aquatic pollutants. transformation and biological effects

AQUATIC POLLUTANTS: Transformation and Biological Effects

Proceedings of the Second International Symposium on Aquatic Pollutants Noordwijkerhout (Amsterdam), The Netherlands September 26-28, 1977

Edited by

O. HUTZINGER University of Amsterdam

and

I.H. VAN LELYVELD and B.C.J. ZOETEMAN National Institute for Water Supply, Leidschendam

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First edition 1978

British Library Cataloguing in Publication Data

International Symposium on Aquatic Pollutants, 2nd, Amsterdam, 1977 Aquatic pollutants. 1. Water - Pollution - Congresses I. Title II. Hutzinger, Otto III . Lelyveld, I H IV.Zoeteman, B C J 628.Γ68 TD420 77-30670

ISBN 0-08-022059-2

In order to make this volume available as economically and as rapidly as possible the authors' typescripts have been reproduced in their original forms. This method unfortunately has its typographical limitations but it is hoped that they in no way distract the reader.

Printed in Great Britain by William Clowes & Sons Limited London, Beccles and Colchester

Preface

Water, like air, is a precious resource which is taken for granted by most people except in areas where this commodity is scarce, as for instance in arid and semi-arid regions.

With increasing population densities in many areas of the world less and less ground-water is becoming available for water supplies and more and more surface water has to be processed for this purpose. This means that municipal water works often have to deal with heavily contaminated raw water since, as with the River Rhine for instance, someone's waste is somebody elsefs resource downstream.

Water contaminated with sewage can, in principle, be treated even when the BOD val-ues are very high. Certain components in municipal and in industrial waste in parti-culer are difficult to remove by most conventional treatment methods. These chemicals, often referred to as micropollutants, are usually anthropogenic (man-made) compounds which are resistant to biodegradation, often bioaccumulate and sometimes possess toxic or other undesirable properties.

The emphasis of the 2nd International Symposium on Aquatic Pollutants was entirely on the compounds described above. The three days of the symposium were roughly divided into

(1) Sources of pollutants, analytical methods, data banks and the natural back-ground of organic compounds;

(2) transport, biodegradation, photodegradation and related transformations; and (3) biological and toxicological effects.

The organizers of this symposium have attempted to attract internationally known specialists for the respective areas to be discussed. We express our gratitude to the scientists who have given their time and who have contributed to the success of this Symposium as speakers and authors.

We hope that this volume will be useful as a review of the present knowledge but also to stimulate further thought and research.

The organizers of the Symposium thank the following organizations for financial assistance: The European Economic Community, the Dutch National Institute for Water Supply and the U.S. National Cancer Institute.

ix

Welcome Speech P. SANTEMA

Director National Institute for Water Supply, Voorburg, The Netherlands

Ladies and Gentlemen, It is a great pleasure to me to welcome all of you, and particularly those coming from abroad, at the 2nd International Symposium on Aquatic Pollutants which is organized by the National Institute for Water Supply and the University of Amster-dam. This 2nd Symposium is a follow-up of a symposium held in April 1974 in Athens, Georgia, U.S.A. on the same topic. North-America en Western Europe, although geographically divided by immense quantities of water, are becoming more and more close neighbours, which is proven once again by this occasion. It is the purpose of this symposium to further strengthen the international collaboration of all those people in science, industry or governmental agencies devoting their activities to control the pollution of the aquatic environment by hazardous chemical compounds.

The Organizing Committee has tried to compose a program covering all environ-mental aspects of chemicals introduced into water. The program includes papers on the origin of chemical pollution, transport and transformation of aquatic pollutants and last but not least their effects on aquatic organisms and man. The program has been focussed on the recent developments in the countries of the European Communi-ties as well as other countries in Western Europe. Furthermore main contributions come from other parts of the world, including the U.S.A., Canada, Israel and Japan.

There are 3 plenary lectures, scheduled at the start of each of the symposium days. These plenary lectures are meant to present a general framework for the detailed scientific papers which will be discussed the coming days.

The first plenary lecture covers the problems in setting sanitation strategies based on toxicological and environmental concepts. After this lecture by Dr. van Esch, the lecture on the second day deals with industry's point of view on pollu-tion abatement, presented by Dr. van Lookeren Campagne, followed on the last day by a survey of european legislation on aquatic contamination by Mr. de Geer. Finally the most controversial items, resulting from the 3 symposium days, can be evaluated in a panel discussion chaired by Mr. van der Veen on Wednesday afternoon.

I am certain that the presence of so many eminent speakers as well as the large interest of about 200 participants will make this symposium a successful one. I am very happy that you all came to Noordwijkerhout even though many of you had to travel over large distances for this purpose. I trust that your participation in this symposium will be sufficiently pleasant and instructive to let you forget for a moment your desire to visit Amsterdam, the windmills and other attractive spots of this country.

Ladies and gentlemen, as it is time to really open this symposium I am very pleased to introduce to you Mr. Reij, Director General for Environmental Protection of the Ministry of Public Health and Environmental Protection of The Netherlands, who will now present his opening address.

XI

Opening Address

W. C. REIJ

Director General for Environmental Protection, Ministry of Health and Environmental Protection, Leidschendam, The Netherlands

Like air, water is transported over large distances. It easily passes national borders and political barriers and in this way also chemical contaminants are carried with the water from one country to another and from man to the organisms living in the fresh and salt water environment. This free circulation of water and its contaminants makes it essential to consider water pollution problems in an international perspective. I am therefore happy that this 2nd International Sympo-sium on Aquatic Pollutants could be organized and particularly that it is held in this country, which, although small, is confronted with severe problems of pollution of the aquatic environment.

As you may know The Netherlands is a country with a good deal of surface water. About 7% of the national surface area consists of water. The North Sea forms the national border to the North and to the West. At the Eastern border the Rhine river enters the country and the smaller rivers Meuse and Scheldt cross the Southern border. In the centre of the country the IJssellake is situated.

However this apparent abundance of water in The Netherlands is misleading. An example in this respect is that Lake Geneva in Switzerland alone contains 18 times as much water as all the Dutch lakes and water courses together. Furthermore water of good quality has become scarce. During the past century the main fresh water source, the River Rhine, has changed from a river yearly supplying the population with many thousands of tons of salmon into a river, now often characterized as the open sewer of Western Europe. Around 1960 the last salmon was reported in this river. Already in 1940 fish from the river was unfit for consumption due to the bad taste.

Nowadays the River Rhine carries the waste water from a population of about 40 million inhabitants of Western Europe as well as the waste water of several of world's largest centres of chemical industry in Basel, Frankfurt, the Ruhr area and Rotterdam. Therefore it is not surprising that at present the water can only be used as a source for drinking water supply at the expense of high treatment costs, that it is unfit for recreational purposes such as swimming and that it poses a potential threat to the aquatic ecosystems in lakes, estuaries and the North Sea.

Besides surface water one might suppose that the rather high frequency of rainy days in this country would result in the availability of large quantities of groundwater. However due to the high population density, the average volume of rain water per head of the population is so low that in this respect The Netherlands is one of the driest areas of the world, after Hong Kong and Singapore. The Netherlands is supplied on the average with only 5 m3 of rain per head per day, in comparison to countries like Israel at 7.5, the United Kingdom at 12, Japan at

Xlll

xiv W. C. Reij

18 and the U.S.A. at nearly 100 m per head per day. This means that the available groundwater in this country is a very precious water source of which the quality has to be carefully protected. For this purpose a special law relating to the pro-tection of the soil against pollution is presently being prepared by the Ministry of Health and Environmental Protection. In the framework of this law a number of preventive regulations have to be specified aiming at, amongst others, protection of groundwater catchment areas for potable water supply. This can be realized by specifying protection zones in which certain potential contaminating activities are not allowed or are bound to strict limitations. Besides preventive actions it has been recently shown that particular attention has to be paid to detection of already existing cases of groundwater pollution by certain hazardous chlorinated chemicals such as trichloroethene. The existence of a relatively widespread con-tamination of our groundwater resources with persistent chemicals once more illus-trates the need to prevent such a pollution, as remedial actions are almost impos-sible or only effective after long periods, contrary to pollution abatement in the case of river water. A further potential source of ground water pollution is the contribution of contaminated rainwater, an area which is now given closer consi-deration than in the past.

Due to the relative scarcity of groundwater, the rivers Rhine and Meuse are of great significance to the aquatic environment and the public water supply of The Netherlands. The high degree of pollution of these industrialized river basins is however in conflict with the different usages of these surface waters, resulting in an urgent need to realize on short term effective remedial activities. As such activities should necessarily have an international basis it is of great significance that recently international agreement has been reached on this problem. After an agreement in principle on a directive of the Council of the European Communities in 1975, relating to the prevention of the discharge of chemical sub-stances into the aquatic environment, a convention against the chemical pollution of the Rhine was signed in December 1976 by the Ministers of the countries concerned. These agreements provide an essential framework for the realization of the sanitation priorities relating to chemicals of the so called black and grey lists.

In order to arrive at a detailed sanitation strategy much additional infor-mation is needed. This information has to be generated by scientists and should include detailed data on origin, transport, transformation and biological effects of chemical pollutants. Here lies the direct connection between the themes of this symposium and the information needed for operational decisions. There exists a close relationship between the availability of scientific data and the realization of sanitation measures. Lack of sufficient information would inevitably result in diminishing progress in improving the quality of the aquatic environment. An example in this respect is the drastic reduction of discharges into the river Rhine of endosulphan and mercury. I am convinced that new data on the behaviour and biological effects of chemicals will significantly contribute to the establishment of a cleaner aquatic environ-ment in the coming years. This symposium offers unique opportunities to evaluate the impact of contaminating chemicals on aquatic organisms and on the health of man, exposed to those chemicals during swimming or by means of consumption of drinking water derived from the contaminated sources. In my opinion quantification of the persistency of chemicals in the aquatic environment and their biological effects on aquatic organisms and man should be the main purposes of research efforts in this area.

Mr. Chairman, ladies and gentlemen, I would like to open this meeting with the hope that the international gathering here will contribute to the advancement of effective procedures needed for the protection and sanitation of the aquatic environment of Western Europe and other parts of the world. I wish you a successful meeting.

Aquatic Pollutants and their Potential Biological Effects

G.J. VAN ESCH

National Institute for Public Health, Bilthoven, The Netherlands

When looking at the map, as a first impression, it seems that the quantity of water on earth is inexhaustible, but only λ% is available as sweet water. From this 1%, 99% is in the bottom and about \% as surface water. The consumption of water by man is increasing and we are faced with an ever-increasing demand for potable water. It will be clear, as was stated at the United Nations Conference on Water Resources held in 1976, that "clean" water is a major need for the pros-perity of the world population now and still more in the near future.

Improvement of our public health, combating of diseases and the high mortali-ty, cannot be realized without the availability of sufficient drinking water and a well-balanced diet and clean air. Clean water is up till now for millions of people living in certain areas of the world self-evident but more than 1500 mil-lions of people do not have the disposal over sufficient potable water. But even in the former areas there will perhaps come a time that there will be a water shortage, quantitively and/or qualitatively.

This is the reason that all over the world research is going on to find ways to purify surface waters and reclaim potable water from municipal waste waters and from effluents.

A big problem in the purification and reclamation of potable water from pol-luted water is the presence of thousands and thousands mainly man-made chemicals that are directly or indirectly released into the environment. Although the con-centrations of these pollutants are still rather low, many of these compounds are toxic to human and/or animal life, some of them are mutagenic and/or carcinogenic or have serious ecological implications. It is therefore necessary to evaluate the potential risks to man and the environment of these chemicals.

IMPACT OF AQUATIC POLLUTION ON THE HEALTH OF M M

In discussing the influence of aquatic pollutants on the health of man, there are three main ways of exposure: dermal contact during recreation (swimming); oral intake of drinking water and the consumption of fish and other waterorganisms.

During bathing and swimming the main contact of man with surface water will be dermal contact. This aspect did up till now not get much attention. In general the concentrations of the pollutants will be low so that it is not likely that these levels will induce irritation of the skin and/or mucous membranes. There are

1

2 G. J. van Esch

a few compounds, such as the well-known dioxins and possibly benzofurans and a few other compounds, that may perhaps induce reactions in certain persons. This aspect should not be ignored completely and attention should be given to the occurrence of these cases.

In principle it can be stated that drinking water has to be as pure as possi-ble. To achieve this a wide range of national and international drinking water quality standards are or will be established. Up till now these standards concern mainly inorganic chemicals and a number of chemical and physical criteria. Quality standards for organic chemicals are mostly still absent.

For the establishment of standards for drinking water, fish and waterorganisms it is necessary to have the disposal of data on the occurrence of these contami-nants in food, drinking water and air. Besides this information, data on shortterm and longterm toxicity, reproduction, teratogenicity, mutagenicity and carcinogeni-city, data on biotransformation (pharmacokinetic studies) and bioaccumulation are necessary. Furthermore it is of great importance to have data on the total body burden and about the presence of the contaminants in human organs and tissues.

For a small number of compounds these data are available but for most of the pollutants occurring in surface water, drinking water and food this is not the case. Because we are dealing with a great number of compounds, to begin with, it is necessary to select those compounds that have the greatest impact on the health of man and his environment. The compounds can be selected on the basis of toxicity, carcinogenicity and/or mutagenicity and/or on the basis of persistence and bioac-cumulation. It is clear that carcinogens (and possibly mutagens) and bioaccumula-ting compounds should have the highest priority. How do we know which compounds bioaccumulate and enter in a foodchain? In our Institute bioaccumulating compounds got attention in the last years. The following study was carried out: Clean cultivated crustaceae (Daphnia magna) were placed in Rhine River water (pho-tograph 1). After a few hours the organisms are collected and a chlorogram is made in comparison with a comparable control group with a mass-spectrometer. At the same time riverwater, fish (eel) and a bird (grebe) was collected and analyzed at the same way. So an impression of bioaccumulation was got in Ma simple foodchain" = riverwater - daphnia - fish - bird. In comparison also a chlorogram was made from human fat (Figs. 1, 2, 3, h and 5).

The grebe and man are both standing near the end of a foodchain, and surprisingly the chlorograms look alike. Besides the well-known chlorinated pesticides as DDT and others, also unknown compounds are present, that have to be identified to find the ultimate source and consequently to take measures of these compounds by legal actions.

Besides the bioaccumulating compounds a great number of compounds is present in surface water, groundwater, reclaimed water and even in drinking water. The levels of these compounds are mostly in the range of a few μg/l and it is still an unanswered question whether drinking water containing a number of these com-pounds have a serious impact on the health of man. To answer this question one could study each individual compound separately on toxicity and calculate an ac-ceptable daily intake as is done for foodadditives and pesticides. But because mixtures of compounds are present the approach is also followed to test extracts of these waters on mutagenicity and in longterm studies to see whether these ex-tracts are toxic, mutagenic and/or carcinogenic.

Another approach is to carry out epidemiological studies with different po-pulation groups that consume different type of waters and estimate whether there is a relationship between mortality, incidence of tumours, heart- and vessel di-seases or other diseases.

All these studies will give information but at the end it is necessary to know which compounds have negativ effects on the health of man, so that steps can be taken to prevent exposure in future.

As already was stated the low levels of these pollutants in drinking water seem to be not of great importance because even for the most toxic pesticides the

Aquatic Pollutants 3

quantities that can be consumed by man, and that are considered to give no harm, are in the order of 100-500 ug/day with his diet. When we consider chloroform,found regularly in drinking water in levels of B few ug up to 300 μg/l, the latter con-centration is still at least a factor of 15 lower than the acceptable daily intake by man per day, calculated on the basis of longterm toxicity studies applying a safety factor of 100. So the intake via water is at least 1500 times lower than the doses that did not give an toxic effect in animals. However, in the case of chloroform the total body burden has to be taken into account, for instance the intake via toothpaste, in which up to h% chloroform is used, is important in con-sidering the impact of chloroform on the health of man. When discussing chloroform we are apart from the toxicity directly involved in the problem of carcinogenicity. A question as: "do there exist no-effect levels for carcinogens", is well-known. But we must realize that this question will not be answered soon. That means that we have to deal with the presence of carcinogens in food and drinking water, with-out having the answer. To protect man and the environment, it can be stated in principle that carcinogens should not be present and that every effort should be made to eliminate these pollutants.

Another important question is "what is a carcinogen"? Also this question is difficult to answer, but there seems to be some evidence that not all compounds that induce tumours are real carcinogens. Real carcinogens induce tumours in dif-ferent animal species in general with low dose-levels. The tumours that are in-duced are mostly of a specific type. Aflatoxins, give livercarcinomas, vinylchlo-rid, angio sarcomas and the different type of nitrosamines induce different type of tumours in specific target organs. The induction of tumours by the other com-pounds can be possibly explained by a. non-specific irritation for example U-ethyl-sulphonylnaphthalene-1-sulfonamide

administrated orally to mice, gives bladderstones, that induce hyperplastic re-actions of the bladderepithelium and consequently bladdertumours occur.

b. hormonal dysfunction, for example high dose levels of an anti-thyroid agent. c. cocarcinogenic or promoting agents. d. microsomal enzyme induction, by certain chlorinated hydrocarbons may have an

negative or positive effect on tumour induction by a carcinogen. e. suppression or overstimulation of the immune system, for example immunosuppres-

sion gives an increase in sensitivity for the induction of tumours. Furthermore the dose level necessary to induce tumours is of importance, it shows the difference in potency of the different carcinogens. Certain carcinogens such as aflatoxin induce already tumours in animals at a daily dose level of a few μgfs, while others have to be given in levels of 50, 100 or even 1000 mg/kg bodyweight/ day (table 1). This aspect has also to be'taken into account, because, as Drückrey has proved, that the daily dose level of a carcinogen and the time of ad-ministration are related to the induction time of the tumours. When low daily dose levels are given, the induction time of the tumours will be long. It even happened that no tumours are induced, the induction time is longer than the lifespan of the animal. With high dose levels the induction time is short. It will be clear that the intake of low levels of a chemical by man, that induce tumours in animals with high to very high daily dose levels, will have a very long latency period to induce tumours and the risk for man will be therefore much lower than for real carcinogens such as aflatoxins and nitrosamines. At last there seems to be evi-dence that the statistical chances that a hit of a carcinogen on a proper region of DNA will be considerably higher with high dose levels than with low dose levels of a carcinogen since the relative modes of detoxification are higher at low dose levels. Furthermore in recent years also so-called DNA-repair enzymes have been discovered, enzymes that detect and remove carcinogen-altered DNA.

All these reasons make it acceptable to believe that the impact from real carcinogens is much more serious than from the other "tumour inducing compounds" and that even for the last group of compounds perhaps no-effect levels exist.

It will be clear that much more research has to be done in the field of car-cinogenesis before we are sure that this suggested difference really exists. How-

4 G. J. van Esch

TABLE 1 DOSE-LEVELS OF HALOGENATED COMPOUNDS IN COMPARISON WITH REAL CARCINOGENS NECESSARY TO INDUCE A (APPROX.) MINIMAL NUMBER OF TUMOURS OR NO TUMOURS

vinylchloride

chloroform

DDT

dieldrin

trichloroethylene

carbontetrachloride

aflatoxin B1

DENA

DMN

AAF

Butteryellow

DMAS

ß-naphthalamine

Benzo(a)pyrene

Saccharine

Animal

species

rat

mice/rat

mice

rat

mice

rat

mice

rat

mice

rat

trout

rat

rat

mice

rat

mice

rat

rat

hamster

rat

mice

rat

rat

Daily dose-level

tumours

9

60

0.5

-

0.05 (?)

TOO (clear effect)

-

1250 (clear effect)

-

0.0001

0.005

0.075

o.k -

2.5

2.5 (?)

0.05 (clear effect)

5 (?)

2.25

1250

in mg/kg b.w.

no tumours

Γ77) 15 (?)

10

-

2.5

500-1000

50-200 (?)

0.0125

0.1 (?)

1.25

0.75 (?)

±50 (?)

+20

500


ever, the meaning of this discussion is to bring the problem in perspective, when dealing with the presence of tumour inducing chemicals in food and drinking water

IMPACT OF AQUATIC POLLUTION ON AQUATIC LIFE

It has been stated by different authorities that "the protection of the health of man starts with the protection of the environment". This means that it will be necessary to decrease the existing pollution and to avoid further pollu-tion. This only can be done by making rules and to avoid pollution and setting standards for pollutants already present in surface water.

From experience it is clear that the standards for waterorganisms will be of a different magnitude as the standards that are established for drinking water and food, in protecting man. As an example can serve endosulfan. On the basis of short-and longterm toxicity data it is possible to establish an acceptable daily intake (ADl)for man of 0.0075 mg/kg bodyweight (CO 0Λ5 mg/person of 60 kg). Endosulfan was present in the Rhine river during a certain period in 1969 and the highest level that was measured in the Netherlands was 0.7 ug/1. From experimen-tal studies with fishes it is known that levels as low as 0.3 ug/1 are toxic to fish. When, hypothetically the surface water of the Rhine River could have been used directly as drinking water (2 1/day) only 1.U ug would have been taken in by man. The difference in sensitivity between man and waterorganism is in this case a factor of about 300 to 750.

Another example may be the presence of cholinesterase inhibiting compounds (expressed as paraoxon). In the Rhine river regular measurements were carried out since 1969. In 1975 a sharp increase was noticed and Daphniae placed in this wa-ter died within hQ hours. The highest level measured was 50 ug/1. Suppose this water had been used as drinking water this would mean an intake of about 100 ug/ day (in the other years it will be only in the order of 5-10 ug/1 or less). The acceptable daily intake for parathion for man is 0.005 mg/kg bodyweight (co 0.30 mg/person of 60 kg). In this case this means a factor of about 3 (in 1975) hut in general this will be 30-60 (Fig. 6 ).

From these and other examples it can be concluded that in general the contri-bution of the pollutant via drinking water will be small in comparison with the intake of contaminants via food and sometimes air by man. In other words when we accept the assumption that human health starts with a healthy environment we have to protect aquatic life that means to set a standard for surface water on this basis. Man will than be automatically protected, of course we must be aware of the fact that perhaps for a few compounds this may be not the case.

How to establish standards for aquatic pollutants in surface water with the purpose to protect waterorganisms?

When we like to protect the ecosystem we have to know what is the impact of a great number of pollutants. This means that these compounds should be tested on toxicity, persistence and bioaccumulation. Because we are dealing with a great number of animal species of different life-stages and of different trophic levels, it is necessary to do studies with, when possible, a number of selected animal species.

The last 5 or 10 years more and more studies are being carried out, never-theless the studies are more or less still in a stage of development. Questions as: "what type of animal species represents the situation in nature"; is it possi-ble to cultivate these animal species in order to have an adequate supply of heal-thy and of uniform size and age the whole year long. Many difficulties have to be overcome, for instance in which stage of development are the animals used, young or adult animals; the problem of administration of the testsubstance; volatile compounds evaporate, other compounds are adsorbed on the glasswall or are broken down. So regularly control of the concentration of the substances to be tested in the water is necessary, what means the availability of sophisticated analytical techniques. Furthermore temperature, oxygen, pH, hardness of the water and other

6 G. J. van Esch

modifying factors can effect the toxicity of the pollutant. Different types of animal species have to be available, because one animal type can be used for acute toxicity studies, others better for embryotoxicity and teratogenicity and/or re-production studies.

In our Institute a number of test animals are available that gives us the possibility to test for short- and longterm toxicity and reproduction; algaea (Chlorella pyrenoidosa), crustaceae (Daphnia magna), fishes (Lebistes reticulatus, Oryzias latipes and Salmo gairdneri). These animals can also be used in bioaccumu-lation studies ("foodchain").

^

Daphnia

Chlorella

ti I

Lebistes

<£* Water

&

^

Daphnia

Lebistes

Even if the experiments with the testorganisms that are used nowadays give certain results it is still difficult to evaluate these results into impact on aquatic life in general. Is it necessary to introduce a "safety factor" for in-stance a factor for the fact that only a few types of animals are used in the tests and that perhaps certain types of organisms are still more sensitive than the most sensitive animal species used in the experiments. Do we like to protect every individual in the aquatic environment, or can it be accepted that a certain percentage of a population will die and the population as a whole will recover?

Before all these questions can be anwered much more research have to be done, however it is not possible to wait with regulations untill all answers have been given.

Several authors have made suggestions to define the maximum concentration at which a chemical may be present in the aquatic environment without damaging its biota. Since all jthese methods have their limitations and are not satisfactory, Canton and Sloofrproposed as a first step in the process of establishing quality standards for aquatic life in surface water an "Ecological Limit" (EL), on the ba-sis of toxicity, persistence and bioaccumulation.

The toxicity of the compound for water organisms of different trophic levels (no effect level) or the lowest concentration at which 10% of the testorganisms in the longterm experiments are affected (=EC Q ) , will be used, assuming that 10% of this effect on one type of organism does not disturb the ecosystem. If the calcu-lation of the EC1 0 is carried out by the method of Litchfield and Wilcoxon (19^9) then the slope of the dose-effect curve is very important. The smaller the slope, the bigger the concentration-range at which 0-100% effect is expected. It was suggested to express this dependency by multiplying the EC with tg Οί (α= slope), when Οί ^. 1|5°. If a > k^° this correction factor will not be used for establishing the E.L.

The persistence and bioaccumulation of the compound are classified arbitrari-ly and for the different groups of pollutants different safety-factors are intro-duced: f(p) for the persistence and f(a) for the accumulation.

* Department Hydrobiological Toxicology, National Institute of Public Health, Bilthoven, The Netherlands

**National Institute for Water Supply, Voorburg, The Netherlands


TABLE 2 A PROPOSAL FOR SAFETY FACTORS (f), BASED ON THE PERSISTENCE (fp) AND THE ACCUMULATION (fa) BEHAVIOUR OF THE TOXICANT

Toxicant

slightly persistent

persistent

very persistent

slightly accumulative

accumulative

very accumulative

Half life value in days

< 10

10-100

> 100

Concentration ratio

< 100

100 - 1000

> 1000

Safety factors (f)

f = 1 P

V 5

f = 10 P f = 1 a f a = 5

f = 10 a I

When we take these factors together the following relation is found:

™ -, · -, τ· -4. fVT\ no-effect level or EC„^ x tg^ (if o:^ k5°) Ecological Limit (EL) = -10 ö — * ——iL—L-p a

The authors are aware of the fact that also other factors such as physical-chemi-cal characteristics of the water and others are of importance, but the above EL is based on several extensive studies and will be applied now in practice to di-vide pollutants in classes and in a later stage to propose standards.

In the framework of the Treaty against chemical pollution of the Rhine river prepared by the International Rhine Committee, pollutants are divided in classes so called black, grey and white lists substances. The principle behind these lists is that black list substances should not be discharged in the river at all, grey list compounds have to be limited in discharge and white list compounds may be discharged without limit.

It will be clear that for many pollutants the toxicological basis to divide them in toxicity classes is not available. Nevertheless when we like to procede and this is really necessary, we have to use the available data and try to classi-fy them on the data available. Canton and Slooff made a scheme in which data on shortterm and longterm toxicity, carcinogenicity, bioaccumulation and persistence are used, in order to classify the pollutants in black, grey and white. It should, however, be kept in mind that the classification of a compound is tentative and has to be revised regularly in the light of new data and new knowledge. The pro-posed classification is as follows:

3 G. J. van Esch

TABLE 3 CLASSIFICATION OF THE COMPOUNDS IN BLACK, GREY AND WHITE LISTS ON THE BASIS OF TOXICITY, PERSISTENCE AND BIOACCUMULATION

Toxicity

acute (LC)t_n in mg/1

< 1

1-10

10-100

> 100

long-term, no-effect level in mg/1 or EC10

< 0.01

0.01-0.1

0.1-1

> 1

Persistence (half-life time) in days

< 10

10-100

> 100

< 10

10-100

> 100

< 10

10-100

> 100

< 10

10-100

> 100

Bioaccumulation (cone.ratio*s)

< 100

black

black

black

grey

grey

black

white

white

grey

white

white

grey

100-1000

black

black

black

grey

black

black

white

grey

grey

white

grey

grey

> 1000

black

black

black

black

black

black

grey

grey

black

grey

grey

grey

The remark should be made that: In using these data it is suggested that when no, or only limited data on toxicity are available, the substance should for the moment be classified as black. When a substance is carcinogenic, or suspicion exists as regards to its carcinogenicity, because of the existence of positive mutagenic data or on the basis of chemical relationship with known carcinogens, these compounds should also be classified as black. If no data on persistence and/or bioaccumulation are available or chemical relationship with known accumulating properties exist, the compounds should be classified as black, except in case the toxicity is low. In the last case the com-pound is grey.

At the moment we have classified approx. 50 substances, as a first start to see whether this approach is useful in practice.

After dividing the compounds in these three classes the next step will be to set standards for the compounds that should be limited the so called grey list com-pounds .

After writing this paper I received the United Nations Environment Programme (UNEP) report (1977) entitled: "Overview in the priority subject area, health of people and of the environment". On page 20 a summary was given about the problem that I have tried to explain to you. I quote:

"The present concern about chemical pollutants in water is not so much with the acute effects on human health (although they do occur), as with the possi-ble long-term effects of low level exposure which are often unspecific and difficult to detect. In addition to the possible effects of ingestion and other direct water contacts, chemicals may influence man's health indirectly by disturbing the aquatic ecosystems or by accumulating in aquatic organisms used as human food. Investigations need to be undertaken on the long-term ef-


fects of the various contaminants with consideration of their chemical and "biochemical transformation which may take place. In the development of water quality criteria, attention should be given to the fact that man may be ex-posed to several toxic water pollutants at the same time and that the effect may he synergistic. Such information is paramount if safe levels are to be specified or standards enacted. Efforts are also required to develop practi-cal guidelines and standards for water quality (drinking, recreation, irri-gation and other uses) which are based on sound scientific criteria."

Photograph 1: Clean cultivated Daphnia magna are placed in Rhine river water to study the bio-accumulation of xenobiotic substancesby these organism.

10 G. J . van Esch

WATER RHINERIVER

10 ppt Cl total weight

Figure 1: Chlorogram of a petroleum-ether ( p . e . ) e x t r a c t of Rhine r i v e r water (Biesbosch) made by gaschromatography-mass-spectrometry (GC-MS).

DAPHNIA MAGNA

NL

0.1 ppm Cl total weight

***** Mm »Aiiwii« ******

Figure 2: Chlorogram of a p.e.-extract of Daphnia magna, placed in Rhine river water (Bies-bosch) for a few hours, made by GC-MS.


EEL

0.1 ppm Cl fat basis

Figure 3: Chlorogram of a p.e.-extract of an Eel, catched in Rhine river water (Biesbosch), made by GC-MS.

DDE

GREBE

1 ppm Cl fat basis

Figure k: Chlorogram of a p.e.-extract of an grebe, that lived in the Biesbosch, made by GC-MS.

12 G. J . van Esch

-uWL-r \A*J

LU

HUMAN FAT

1 ppm Cl fat basis

Figure 5: Chlorogram of a p.e.-extract of human fat, made "by GC-MS.

Cholinesterase inhibiting substances in the River Rhine (Lobith) concentration yug/l (as Paraoxon) 50r

1976 1977 I I

Figure 6: Occurrence of cholinesterase inhibiting substances in the Rhine r iver in the period of 1969-19TT.

Chemicals with Pollution Potential

O. HUTZINGER*, M. Th. M. TULP* and V. ZITKO**

* Laboratory of Environmental and Toxicological Chemistry, University of Amsterdam, Nieuwe Achtergracht 166, The Netherlands

^Environment Canada, Biological Station, St. Andrews, N.B., Canada

ABSTRACT

The pollution potential of a chemical compound depends on the intrinsic properties which are determined by its structure on the one hand and a number of non-chemical factors related to production, transport and use pattern on the other. Generally speaking, a chemical compound is more likely to be a serious pollutant if it fulfils most, if not all, of the following criteria: large (industrial) production, use which makes environmental leakage likely, high dispersion tendency, pronounced per-sistence, tendency to bioaccumulate and high toxicity. All these criteria relate to (i) the intrinsic toxic property of the chemical and (ii) three rate factors re-lated to behaviour in the environment (ecokinetics) i.e. rate of release into the environment, rate of disappearance from the environment and rate at which the com-pound or its degradation product becomes available to an organism in question. Environmental behaviour and toxicity of chemicals can be investigated by laboratory tests (e.g. toxicity or bioaccumulation tests), the benchmark approach (prediction of environmental properties from physical parameters), structure-activity relation-ships, mathematical models or a combination of these. For illustration several examples from the authors' laboratories are given in the areas of bioaccumulation, metabolism and toxicity using series of haloaromatic compounds as substrates

INTRODUCTION

Considering the number of chemical compounds which are known and recorded in Chemi-cal Abstracts (> 3 million) the number of chemicals which have become serious pol-lutants is very small. Many compounds described are only prepared in small quanti-ties but even compared to the number of chemicals produced industrially in signifi-cant quantities (> 10.000 compounds produced in quantities larger than 500 kg) the problem compounds are an unimportant number. With respect to this relationship two questions can be asked:

1. What makes a chemical compound a pollutant; What features make some chemicals real or potential environmental hazards and others not? and

2. Are there perhaps many more compounds with pollution potential but do they, as yet, go unnoticed?

The first question in its theoretical implication has been asked by environmental scientists many times in the past. The second question is recently being asked by

13

14 0. Hutzinger, M. Th. M. Tulp and V. Zitko

various governments of technologically advanced countries and has led to unpreceden-ted regulatory activity manifested in the different "toxic substances laws" (1). The vast amount of knowledge on the pollution potential (environmental behaviour and toxicity) of chemicals required for regulation brings the first question on the features responsible for making a chemical a pollutant out of the theoretical-academic interest into the practical sphere and thus requires a pragmatic approach for fast and efficient accumulation of relevant data. With regard to the expected regulatory activity the following a priori statements can be made: 1. CHEMICALS WILL BE REGULATED (a decision on "what is a pollutant"

will be made), 2. THE LEGISLATOR WILL DECIDE based on a_. information from scientists

b_. prevailing values of society (public opinion) , 3. DECISIONS WILL ESSENTIALLY BE MADE ON INSUFFICIENT EVIDENCE.

The responsibility of the scientific community at this time is to provide the legis-lator with as much factual information and with as many sound concepts as possible to ensure that rational decisions can be made. Before we offer such advice, however, we must be clear in our own minds about the underlying concepts about what makes a chemical a pollutant. The rest of this paper is divided in three sections in which an attempt will be made to: (i) outline criteria for "what makes a chemical a pollutant", (ii) provide in-formation how these criteria can be measured and predicted for compounds and groups of compounds and (iii) illustrate some of the points by providing data and examples from our own work. Before going to the next section a brief historical overview of development of philosophies and approaches in water pollution will be considered.

1. React only after the presence of a pollutant in the environment has been recognized (e.g. Hg, PCB, DDT).

2. Monitor effluents for possible pollutants (e.g. GC-MS analyses). 3. Study possible pollutants (e.g. industrial chemicals). 4. Study relationship between chemical structure and "pollutanf'-properties.

While point 1 reflects an attitude of the past, number 2 is a useful approach which is practised widely since actual pollution problems can thus be found and corrected. The concept in point 3 is that of the toxic substances laws: any compound produced may enter the environment and environmental properties (pollution potential) should therefore be known. Point 4 is a hope for the future which may result in the design of environmentally safe chemicals.

WHAT FEATURES MAKE A CHEMICAL COMPOUND A POLLUTANT (CRITERIA)

Pollutant chemicals may have a multitude of undesirable effects, but in the vast majority of all instances this will be a biological effect. This neglects the small number of cases where pollutants interfere with the physical environment such as, for instance, carbon dioxide with the heat balance of the earth. There is common agreement that a synthetic industrial chemical compound is more likely to be a serious pollutant if it fulfils most, if not all, of the following criteria (3-5): large industrial production,

use which makes environmental leakage likely, persistence, bioaccumulation, toxicity.

With some corrections these criteria can be considered a good set of features for characterizing a compound as a water pollutant. The corrections are mainly concern-ing the fact that natural products can be pollutants as well and that other un-desirable biological effects except toxicity have to be considered also. One compound may serve as an example for both points. The earthy-smelling compound geosmin is produced by microbiological action in storage reservoirs of water works (it is thus

Chemicals with Pollution Potential 15

a natural product), its presence in water is undesirable not because of its toxi-cological properties but because of the unpleasant taste and odour of water con-taining this contaminant (2). In order to appreciate the philosophy of the above-mentioned criteria one may begin with the most simple possible relationship in pollution: release of chemical com-pound on the one hand and biological effect on the other (Fig. 1).

CHEMICAL

COMPOUND

BIOLOGICAL

EFFECT

release contact with organism

Fig. 1 Relationship between a chemical compound and the biological effect in an exposed organism.

Two factors are important in this relationship (i) the nature and severity of the biological (toxic) effect and (ii) how much of the chemical is available for the organism. Such a relationship is straightforward, for example, for a chemical used as drug or a chemical in occupational exposure where effect on the organism (e.g. man) depends on the intrinsic toxicological properties of the chemical and the to-tal amount available to the organism (the dose). With environmental contaminants the situation is more complicated because the en-vironment is "between" the point of release and the point of contact i.e. these two events are separated in time and space (Fig. 2).

Release of

chemical ENVIRONMENT Organism target

site -♦► toxicity

environmental chemodynami c s (ecokinetics)

pharmaco-kinetics

Fig. 2 Relationship between a chemical compound and the biological effect in an organism exposed via the environment.

It is well-known to pharmacologists that the total amount of toxicant available at the target, although determined by the dose, depends on uptake, elimination and metabolism of the drug i.e. its pharmacokirietic properties. It has been suggested to consider the environment as complex super-organism featuring similar properties of "ecokinetics" or "environmental chemodynamics". This concept is shown, using simplified rate constants for illustration in Fig. 3. Once released into the environment at a specific rate (kr) the compound may be transported and may accumulate to make the compound (toxicant) available to any given organism at a specific rate (k-j-) . Factors such as sedimentation and chemical and biochemical degradation remove the chemical competitively (ks) into what is often called a sink. The amount of compound (toxicant) which becomes available to the organism (the dose) thus depends on the rate of release, the rate of disappear-ance in sinks and factors such as transport and biöaccumulation. A complicating factor arises through formation of metabolites or other decomposition products which are pollutants in their own right (km; e.g. DDT—►DDE).


release

ENVIRONMENT

-► transport ι accumulation

sedimentation

degradation

♦-organism

SINK

Fig. 3 Representation of environmental chemodynamics (ecokinetics)

kr rate of release of chemical rate at which chemical (toxicant) becomes available to organism same for metabolite rate of disappearance of chemical (sink)

For proper modelling the simplified general environmental scheme of Fig. 3 will, of course, have to be divided in submodels e.g. in a lake-, river-, estuary- etc. submodel with application of proper and detailed rate constants for each compound. There is now a considerable body of scientific literature on the environmental fate of chemical compounds (e.g. 6-9).

Table 1 gives a systematic summary of factors responsible for the pollution poten-tial of a chemical compound. The three most important ones which can be investiga-ted by scientific methods are: bioaccumulation, persistence and chronic toxicity. Some methods for determining or predicting important environmental features are described in the following section.

TABLE 1 Important Factors Responsible for Pollution Potential of Chemical

FACTORS IMPORTANT FOR POLLUTION POTENTIAL OF CHEMICAL question asked depends on

1 ENVIRONMENTAL RELEASE (entry of chemical into environment)

how much enters the environment

total amounts produced, pro-duction techni-ques , shipment-, use and disposal pattern, involun-tary production (natural product)

2 ENVIRONMENTAL DYNAMICS (ecokinetics)

how much remains available to various organisms and in what concentration

dispersion beha-viour , bioaccumu-lation, persistence availability of sinkd

3 BIOLOGICAL EFFECT (toxicity)

nature and severity of toxic effect

pharmacokinetics mode of action

variable parameters dependent on "human" factor

unchangeable properties determined by chemical structure

kr kr kr


HOW CAN "POLLUTION FEATURES" (CRITERIA) OF CHEMICALS BE MEASURED AND PREDICTED

The need for sound, reliable data which can be obtained reasonably easily was dis-cussed in the introduction. Obviously, not all chemical compounds can be investiga-ted for pollutant properties and effects. The first selection of classes of compounds will be on large quantity products and compounds with known or suspected undesirable properties (toxicity, persistence) or structural relatives of such compounds (5,10). The choice of chemicals for environmental effect studies was discussed recently (5) and appropriate literature consisting mainly of government reports is cited in this publication. The major toxic substances laws have candidate lists of compounds to be investigated (1). "Grey" and "black" lists describing undesirable compounds have been published (11) and a recent MARC-report (12) outlines a detailed plan for ha-zard evaluation. Production and use data for industrial chemicals will be disclosed under the current environmental legislation. Some information is now available from government reports (e.g. US Tariff Commission Report 1972 and 1975; cf. 5). Generally, compounds which are produced in quantities of about 25 x 10° kg/year or more and persist in the environment for more than 3 months have the potential of be-coming widespread environmental contaminants.

The environmental effect evaluation of chemicals i.e. the determination and predic-tion of their environmental fate and toxicological behaviour is possible by a number of different methods. These procedures which are outlined in Table 2 can be used alone or in combination.

TABLE 2 Methods for Measurement and Prediction of Environmental and Toxicological Behaviour of Chemicals

1. LABORATORY TESTS e.g. toxicity, bioaccumulation, biodegradation, model ecosystem.

2. BENCHMARK APPROACH use of physical parameters to predict environ-mental behaviour e.g. volatility - air transport, n-octanol / water partition coefficient - bio-accumulation .

3. STRUCTURE-ACTIVITY RELATIONS e.g. Hansch QSAR. 4. MATHEMATICAL MODELS e.g. atmospheric distribution. 5. PREDICTIVE INTEGRATED RATING SYSTEMS e.g. Crosby or Weber pesticide

hazard rating

1. Laboratory tests. This approach is well-known from the classical LD50 and LC50 tests, carcinogenicity tests etc. The new Japanese chemical substances law, for instance, requires three types of tests, i_ a biodegradation test with sewage sludge; ii a bioaccumulation test using certain fish species and iii a series of tests for chronic toxicity (13). Evaluation of chemicals is possible with tests which are more complex and integrated e.g. the microcosm (model ecosystem) approach (14). As an example of the tests required for carcinogenicity hazard assessment, the pro-posed EPA outline (15) is listed in Fig. 4.


ALL CHEMICALS IN RISK CLASSES

J TIER - 0

CHEMICAL CHARACTERIZATION AND

PERSISTENCE TESTING

TIER - 1

BIOLOGICAL ACTIVITY TESTING

T POSITIVE TIER - 2

PRE-CHRONIC TOXICOLOGY

TIER - 3

LONG TERM BIOASSAY ONCOGENICITY

Fig. 4 Simplified hierarchical testing scheme proposed by EPA

CARCINOGENESIS OF CHEMICALS IN RISK CLASSES.

The proposed tests of TIER - 0 to TIER - 3 are as follows : IMPORTANT TESTS IN TIER - 0 CHEMICAL CHARACTERIZATION. Solubility, ionization constants, adsorption, partition coefficients. PERSISTENCE TESTING Photochemical : activity, sensitivity Chemical : reductive, hydrolytic, oxidative Microbial : aerobic, anaerobic

BIOLOGICAL ACTIVITY TESTING TIER - 1. MINIMUM TEST BATTERY One bacterial test each for gene mutations and for primary DNA damage. One eucaryotic microorganism assay for primary DNA damage. One mammalian cell culture assay for gene mutations. One hamster embryo cell oncogenetic transformation assay or mouse fibroblast oncogenetic transformation assay. One human cell culture assay for inhibited DNA repair or unscheduled DNA synthesis.

ALL TESTS WILL REQUIRE : Confirmation of cellular viability under all test conditions, activated and unacti-vated test systems, positive and negative controls from the same subclass.

PRE-CHRONIC TOXICOLOGY TIER - 2. ACUTE TOXICITY Lethality estimate - 14 day observation period, minimum of 5 dose levels, two rodent species/ both sexes.


REPEATED DOSE STUDY LD - 14 day treatment maximum, minimum of 10 dose levels, two rodent species, both sexes. SUB CHRONIC TOXICITY Estimate of MTD based on 90 - day treatment maximum. Morbidity and mortality observation for 90 days. Minimum of 5 dose levels, spaced to include estimated MTD and 10 MTD. Two rodent species, both sexes.

LONG - TERM BIOASSAY FOR ONCOGENECITY TIER - 3. Life time study with technical grade testing material : two rodent species, both sexes. Morbidity and mortality observations, histopathologic examinations with light and electron microscopy. 5 Dose levels minimum, covering the range including MTD and 10 MTD.

Biodegradation, an important feature in determining the pollution potential of a compound, can be measured in the laboratory by determining the half life of a chemical when treated with sewage sludge or soil microorganisms. A reasonable scale for biodegradation categories has recently been suggested (16,17; Table 3).

TABLE 3 Biodegradability Categories.

Cate-gory

1

2

3

4

5

Biodegradability

Easily degraded

Degraded without much difficulty

Difficult to degrade

Very difficult to degrade

Refractory

Persistence in unadapted soil

1-3 weeks

1-3 months

3 months -1 year

1-2 years

longer than 2 years

Success of biological treatment

Susceptible to normal waste treatment

Susceptible to normal waste treatment

Prolonged treat-ment needed

Leakage possible even with pro-longed treatment

Cannot be trea-ted biologically

Typical chemical

Acetic acid

Benzoic acid

ε-Caprolactam

Chlorobenzene

Hexachlorobenzene

In general, laboratory tests which are designed to answer specific questions when experimentally well planned and executed are the best choice for predicting a com-pound's behaviour in a real situation. The disadvantage of these tests is that only one compound can be tested at one time and that these tests are usually ex-pensive and time-consuming. Furthermore, even with carefully chosen experiments, laboratory data may still inaccurately describe the real situation (e.g. biode-gradation, sludge-environment, LD50 rat-toxicity humans).


2. The benchmark approach. Here an attempt is made to predict environmental beha-viour from physical parameters. The following relations have been suggested :

TABLE 4 Examples for the Benchmark Approach

Physical parameter H' +- Environmental to be measured behaviour

vapour pressure evaporation rates solubilities

soil column percolation adsorbent coated tic adsorption isotherms

exposure at different pH

exposure to reactive species (e.g. oxidants)

irradiations under different conditions

n-octanol / water partition coefficient

transport, dispersion tendency

adsorption sedimentation leaching

chemical stabilities

atmospheric behaviour

photodegradation

bioaccumulation

An illustration of this approach is shown in Fig. 5. It was found that bioaccumu-lation behaviour for a series of compounds can be predicted from the n-octanol / water partition coefficient (18 - 21). Experimental data for both bioaccumulation from a laboratory exposure and partition coefficients for a set of compounds are shown in Fig.5.

fc8

7H

S6

2 5· C (L· U c o u o in

°, 1

A-l 3

2

A = p-dichlorobenzene

B = biphenyl

C = 2-biphenylphenyl ether

D = hexachlorobenzene

E = 2,2',4,4'-tetrachlorodiphenyl ether

i. -B

-T T 1 1 1 1 1 1 1

1 2 3 4 5 6 7 8 9 Log Partition Coefficient

Fig.5. Relation of partition coefficient and bioconcentration of selected chemicals in trout muscle (adapted from 18)

to predict


The bioaccumulation factor can directly be read from the graph for any new com-pound for which the partition coefficient can be measured or calculated. The pos-sible advantage of this approach is that after a relationship has been determined new compounds can be fitted in reasonably easy by determining one or several phy-sical parameters.

3.Structure - activity relations. Quantitative relationships between structure and properties of organic molecules on the one hand and biological activity on the other (quantitative structure activity relationships, QSAR) are frequently deter-mined and have shown to be useful in the development of drugs for many years. A similar approach is now being used in the development of pesticides. The selec-tion of a pesticide for example, is preceded by screening of the activity of a number of analogs and isomers and the establishment of structure-activity r_eJLa.fciP.n-ships for target and non-target species (i.e., an insect and a small mammal). Quantitative structure activity relationships are much less commonly used to pre-dict environmental properties and undesirable biological effects of chemicals ge-nerally. There is now enough evidence (22) to indicate that QSAR are very useful for predicting environmental properties of chemicals and specific toxicological effects. The main difficulty with predicting undesirable biological effects is, that unlike in pharmacology where the activity to be investigated is well defined and accurately measureable, the biological parameters to be considered for OASR are not known beforehand (chronic toxicity ?, LC n ?, cholinesterase inhibition ?, carcinogenicity ? etc.).

Of the many approaches in QSAR the Hansch model (23,24) is the best-known. The con-stants in this relationship are related to the structure and characterize the hy-drophobic, electronic and steric properties of molecules. Relationships have the general form given in the equation below :

log i = k(1).(log P ) 2 + k

(2).log P + k(3).a + k

(4).E + k(5)

c s

where c = molar concentration, characteristic for certain activity, P = n-octanol / water partition coefficient, characteristic of

hydrophobic (lipophilic) properties, σ = Hammett constant, a characteristic of electronic effects.

E = Taft constant, a characteristic of steric effects,

k = constants obtained by fitting the equation to experimental data.

The rationale behind the equation and some of its applications were recently re-viewed by Hansch (24). Other than σ and E , additional so-called free-energy con-stants may be used (25,26) and the introduction of molecular symmetry-related con-stants may be worthwhile since, in general, increasing molecular symmetry leads to higher toxicity (27) . Partition coefficients of a large number of compounds were summarized by Leo et al. (28), and the data can be used to estimate partition coefficients of unlisted com-pounds. A number of π values is also presented in this reference. Additional π values are given in (23) and numerous free-energy related constants in (26) . More recently developed relationships such as the additive model of Free and Wilson, and quantum chemical models have been used in drug research. These techniques have not yet been applied to aquatic toxicology, and the reader is referred to a review by Redl et al. (29) for details. This review also mentions advanced data-fitting techniques based on pattern recognition, such as cluster analysis. A method based on the summation of "hydrophobic fragmental constants" has been de-scribed (30,31) and usually gives log P values in agreement with those obtained experimentally or from the Hansch model (28,29). New structure activity relation-ships usually hold well within related series of compounds. This means that only classes and not individual compounds need to be investigated.


4. Mathematical models. More or less sophisticated mathematical models are some-times used to predict the complex behaviour of pollutants in the environment. While prediction of certain isolated parameters as for instance photochemical behaviour is now possible (32) , there seems to be agreement that in most instances mathematical models are not yet predictive for complex interrelations and in lar-ger environmental sections (33).

5. Predictive integrated rating systems. These systems attempt to rate quantitati-vely, based on the factors outlined in this paper, environmental hazards of chemi-cals. There is a general agreement on the factors, but the outcome of the rating depends on the weight the systems assign to the individual factors. The results may be strikingly different. For example Crosby (34) defines the hazard of a pes-ticide (H) as :

where P

50

= proportion of a compound in organic phase in equilibrium with water, a measure of bioconcentration potential,

= proportion of starting material remaining after 48-h exposure in water or aqueous alcohol to > 300 nm UV light, a measure of chemical stability,

= partition coefficient between organic matter and water, a measure of the tendency to be removed from water by adsorption, and

= median immobilisation concentration for Daphnia Magna.

Crosby suggests that pesticides with H > 100 deserve a careful scrutiny. Weber (35) rates pesticides by considering 1_. acute toxicity (LD-n) to rats, 2_. acute toxicity (LC ) to fish, 3_. longevity in half-life ("converted to full-life"?) and 4_. bioaccumulation factor. Each of these is rated on 1-low, 4-high scale and the values are added so that the overall hazard is expressed on a 4-low, 16-high scale, i.e. 1+1+1+1 = 4 is the least dangerous and 4+4+4+4 = 16 the most dangerous compound. An extensive fractional rating scale (in 0.1 steps) for toxici-ty, longevity and bioaccumulation is given in (35) and a short version is shown in Table 5.

TABLE 5.

Rating scale

1-1.5

1.5-2.5

2.5-3.5

3.5-4

Pesticide Rating Scales

1

mg/kg

2 LC fish ppm

> 4000 > 100 relatively nonhazardous

300-4000 1-100 slightly hazardous

20-300 0.01-1.0 moderately hazardous

< 20 < 0.01 hazardous

of Weber (35).

3 Longevity soil life in weeks

< 15 readily degradable

15-45 moderately degradable

45-75 slowly degradable

> 75 persistent

4 Bioaccumulation factor 1 cone, organism/ c.water

< 60 nonaccumulative

60-700 slightly accumulative

700-8000 moderately accumulative

> 8000 highly accumulative

LD rat

R

om Κ

I


A comparison of the two hazard assessment systems is interesting (Table 6) :

TABLE 6 A Comparison of the Hazard Rating Systems of Crosby (34) and Weber (35).

Pesticide

Parathion

DDT

Lindane

2,4 - D

Crosby

14,000

620

105

6.5

Hazard

Weber

9.2

14.2

11.6

5.9

Both systems assign a relatively low hazard to 2,4 - D but differ considerably in rating parathion and DDT. This difference illustrates the effect of criteria se-lection on the outcome of the rating. Organophosphate pesticides are generally more toxic to Daphnia than to fish, and this contributes to the high hazard of pa-rathion in Crosby's scale. At the same time, parathion is less persistent than DDT thus lowering its rating on Weber's scale. Crosby is using P as an indicator of bioaccumulation potential, whereas Weper is using bioaccumulation factor (concen-tration in organism/concentration in water). Crosby's P or the more generally used octanol/water partition coefficient are measures of uptake regardless of the sub-sequent fate of the compound on the organism. On the other hand, bioaccumulation factors are calculated from tissue concentrations of the administered compound and possibly its metabolites. Compounds taken up and decomposed rapidly in the orga-nism may have zero bioaccumulation factors in spite of high partition coefficients. Obviously, both partition coefficients and bioaccumulation factors have to be con-sidered in hazard assessments. Quantitative and even qualitative rating of hazards of chemicals is a complex and difficult process, with judgement and opinion play-ing significant roles in the selection and weighing of criteria. A general study of the applicability of OSAR to toxicity predictions was carried out by Craig and Waite (36). They reviewed the availability of toxicity and chemical properties data and techniques for deriving QSAR, such as the Hansch multiple regression, the Cramer method of predicting toxicity of a compound from toxicities of its frag-ments, discriminant analysis, and pattern recognition.

The octanol/water partition coefficient appeared to be the most useful parameter for toxicity predictions. A modified Cramer method, based on a learning set of 800 compounds was used to predict the toxicity (rat LDcr») o f ^1 at random selected chemicals. Of these, no predictions were possible for 8 due to lack of appropriate structures in the learning set. The toxicity was predicted correctly within one order of magnitude for 17 compounds. This is a reasonable degree of success, con-sidering the random selection of chemicals and toxicities ranging over 6 orders of magnitude. Pattern recognition techniques did offer a particular advantage, but the discriminant analysis was useful for the preliminary identification of highly toxic chemicals. The structural keys forming the basis of the discriminant analy-sis are presented in Table 7.

TABLE 7. Structural Keys for Discriminant Analysis

Phosphorodithioate 2,3-Dihydrobenzofuran Cyclopentyl ring Piperidine C=N attached to ring Cyclopropyl Isochroman Ring-methyl

Acetylenic Primary amide Aliphatic halogen Trialkyl phosphate

The presence of one or several of these structures may be indicative of high toxi-city. The original report should be consulted for additional details.


CORRELATION STRUCTURE - ENVIRONMENTAL PROPERTIES AND TOXICITY

SOME EXPERIMENTAL DATA

Structure - Toxicity in Aquatic Organisms.

Our studies on the lethality of selected aromatic compounds to juvenile Atlantic salmon (Salmo salar) (37) indicate that the octanol/water partition coefficient and the presence and number of functional groups, such as -OH, -COOH, -SO-H and -NH9, are the main factors determining the lethality of aromatic compounds to sal-mon (Fig.6).

log

7-1 LC

6 1

5 H

4 H

LC mg/l of M=150

h 0.015

I I I I

VERY TOXIC

L C < 1 m g / l

log P

h0.15

M . 5 -

1 5 TOXIC

LC 1-100 mg/ l

PRACTICALLY Ν0Ν-TOXIC

XA LC> 100 mg/l

150·

Fig.6.Lethality of aromatic compounds to juvenile Atlantic salmon. LC = lethal threshold, mole/1 P = octanol/water partition coefficient

1 aniline 2 phenol 3 benzoic acid 4 2-chlorobenzoic acid 5 salicylic acid 6 4-t-butylbenzoic acid 7 catechol 8 1,4-dihydroxybenzene

9 2,6-dinitroaniline 10 4-chloroaniline 11 2-chloroaniline 12 2-aminonaphthalene 13 3-carboxy-B-naphthol 14 3,4-dichloroaniline 15 2,5-dichlorophenol 16 3,5-dichloroaniline

17 3-naphthol 18 2,6-dichloroaniline 19 2,5-dichloroaniline 20 a-naphthol 21 3,5-dichlorophenol 22 pentachloropyridine

The presence of sulfo groups results in practically non-lethal compounds. Simple aromatic compounds such as aniline (1), phenol (2), benzoic acid (3) and its deri-vatives appear to form one group, whilst substituted phenols, anilines and some naphthalene derivatives appear to form another group (see equations I and II )

log ( ~ ) = 0.26 log P + 3.13 ( I )

log ( j -) = 0.27 log P + 4.34 ( ID j


There is an indication that the lethality of naphthalene derivatives follows equa-tion III rather than equation II.

log (i~) = 0.89 log P + 2.54 (III)

With only two exceptions so far the positions of substituents in di- and higher substituted benzenes and naphthalenes change the lethality of the isomers by less than one order of magnitude. One exception is resorcinol, which is non-lethal at the highest concentration tested, 30 mg/£, and this can be related to the well-known differences in the redox potential of 1,2-, 1,4- and 1,3-dihydroxybenzenes. The other exception is 2,7-dihydroxynaphthalene. Both equations II and III predict reasonably well the mean lethality of these compounds (estimated log P = 2.54).

Dihydroxynaphthalene Log (—) , LC in mole/Ä

1,5-1,7-2 , 3 -2 , 6 -2 , 7 -

4.27 4.66 4.66 5.01 5.35

The lethality of pentachloropyridine (22) appears to fit the equation II, but this is only coincidental since the dichloropyridines tested (2,3-, 2,6- and 3,5- esti-mated log P = 1.92) had

log (—) > 3.69. Tri- and tetrachloropyridines were not available. Their expected

log (—)'s are 3.90-4.15 and 4.55-4.65, respectively, based on the estimated log P's

of 2.49 and 3.06.

Structure - Bioaccumulation; Some Inconsistencies in Bromo- vs. Chloro Aromatic Compounds

From the literature and our experiences QSAR is a useful tool for the studies of environmental properties of chemicals, but must be applied with caution, since many, not immediately obvious factors may play a role. A comparison of chlorinated and brominated aromatic compounds serves as an example. The substitution of chlorine by bromine increases the octanol/water partition coefficient (π = 0.71 and 0.86, respectively). Consequently, bromobenzenes and bromobiphenyls should be taken up and accumulated by aquatic fauna to a higher degree than the corresponding chloro compounds. Experiments show that bromobiphenyls with up to 4 bromine atoms resemble quite closely the corresponding chlorobiphenyIs (38). On the other hand, the behaviour of brominated biphenyls with six or more bromine atoms per molecule is quite dif-ferent (39). Hexabromobiphenyls accumulate more slowly than hexachlorobiphenyls. Hepta-, octa- and nonabromobiphenyls are not accumulated from water and to a very small extent from food. In the latter case apparently a hexabromobiphenyl is formed by partial reductive debromination (Fig. 7). Higher molecular weight and lower stability of C - Br as compared to C - C£ bonds (68 and 81 kcal/mole, respectively) may be some of the reasons for these differen-ces, which are not limited to substituted biphenyls. Contrary to hexachlorobenzene which is strongly accumulated in fish from both water and food exposure, hexabromo-benzene is not accumulated at all from either route (38). If increased reactivity of highly brominated benzenes and biphenyls is the reason for the differences be-tween these compounds and their chloro-analogues, then there may be differences in their toxicological behaviour as well. For example, the formation of covalent bonds to some tissue components would be likely. In addition, the relatively low accumulation of highly brominated benzenes and biphenyls does not provide the "early warning system" of environmental contamination, based on the analysis of biological samples.


Br.

100 150 Scon Number

Fig. 7 Reconstructed gas chromatograms of a commercial octa-bromobiphenyl (OB) preparation (A) and extract of fish fed OB-contaminated food (B). Brn, CJln = biphenyls with n bromine or chlorine atoms,

respectively. Chlorobiphenyls and DDE are background contaminants in the fish.

Structure - Bioaccumulation: Measurement of Bioaccumulation by an Internal Standard Technique and Importance of Factors Other than Lipophilicity.

The determination of bioaccumulation of a chemical compound in an organism, for instance from water, requires complete analysis of the compound in two different matrixes (e.g. fish and water) including all standardisation procedures. By the choice of appropriate internal reference standards which are administered in about equimolar concentrations and which have to be well separable by gas chromatography, relative bioaccumulation values with respect to the internal standards can be ob-tained for new compounds. Extraction efficiencies and analytical procedures for the internal test series has to be worked out but once and provided the bioaccumulation of the new compound falls within the series and extraction efficiencies can be assumed to be similar to the test series, reasonably accurate relative bioaccumula-tion values can be obtained by this fast and simple method.

Figure 8 shows a GC tracing of an equimolar quantity of six chlorobiphenyls in which one ring is substituted with two chlorine atoms in the 2,5-position and the second ring of the biphenyl system is substituted from zero to five chlorine atoms respectively. The test compound 2,5-dichloro-4'-isopropylbiphenyl, a representative for the PCB replacement mixture chloroalkylene (40), is well separated from the six chlorobiphenyls (upper trace). The lower trace shows purified extract from abdomi-nal fat of rats to which the test mixture in peanut oil was administered by stomach


tube two weeks before the animals were sacrificed (41). Data on the structural fea-tures in substituted biphenyls responsible for the retention in abdominal fat of rats after administration as a single dose are shown in Fig. 9.

U 10 time (min) 15 20

STANDARD

Li^ji^jLiL· VI V__

10 time (min 20

FAT EXTRACT

Fig. 8 GC tracing (total ion chromatogram) of an equimolar mixture of 6 chlorobiphenyls and 2,5-dichloro-4'-isopropylbiphenyl. Lower trace shows compounds from purified abdominal fat extract. Note the difference of height of peaks corresponding to the test compounds. Unmarked peaks correspond to coextracted products from the fat.

28 0. Hutzinger , M. Th. M Tulp and V. Zitko

LIPOPHILICITY STERIC METABOLISM EFFECTS

Fig. 9 Relative retention of substituted biphenyls in abdominal fat of rats two weeks after single dose administration by stomach tube. Values are relative to the compound which is retained in highest amounts (= 100). Calculated log P values indicated at base-line.

The data show that lipophilicity (expressed as n_-octanol/water partition coeffi-cient) is an important factor in the retention of these compounds in the abdominal fat of rats, but not the only one. From the left part of Fig. 9 it is evident that equimolar doses are retained to a different extent. Generally the retention increa-ses with increasing lipophilicity, however, the retention of the isopropyl compound and the (sterically most hindered) heptachlorobiphenyl does not follow the log P value and the relation of the others is not linear. The center part of the figure shows the relative retention of a series of symmetrical hexachlorobiphenyls with the same calculated log P. The different retention of these isomeric compounds may be due to a number of factors including different rates of metabolism and increased steric hindrance due to chlorine substitution in the positions ortho to the phenyl-phenyl bond. The right part of the figure compares two chlorobiphenyls with alkyl-chlorobiphenyls of approximately the same calculated log P value. The lower reten-tion of the alkylchlorobiphenyls is most likely due to the more facile metabolism of these compounds compared to corresponding chlorobiphenyls. Alkyl-side chain oxydation is indeed an important metabolic route for chloroisopropylbiphenyls in a number of species (41, 42; Fig. 10) and of 4-isopropylbiphenyl in different mammals (43). While these data on the factors responsible for bioaccumulation of chlorobiphenyls and alkylchlorobiphenyls in abdominal fat from rats can only be considered prelimi-nary, they nevertheless strongly indicate the importance of factors other than par-tition coefficients in these series of compounds.


H3SiCH3 H3%*CH2 H 3 C V ° H "^'°

OH

Fig. 10 Metabolism of 4-chloro-4'-isopropylbiphenyl by a mixed culture of aerobic bacteria from activated sludge. The chloroisopropylbiphenyl was used as the sole carbon source (from Ref. 42).

ACKNOWLEDGEMENTS

The second author gratefully acknowledges a fellowship by the Organisation for the Advancement of Pure Research (Stichting Z.W.O., the Netherlands). We are indebted to J. van der Born for valuable assistance with the experiments.

REFERENCES

Examples of "toxic substances laws" in different countries: Switzerland (1969), Bundesgesetz über den Verkehr mit Giften. Japan (1973), Chemical Substances Law. Sweden (1973), Act on products hazardous to man or the environment. Great Britain (1974), Control of Pollution Act, and Health and Safety at Work Act. Canada (1975), Environment Contaminants Act. Norway (1976), Act concerning the control of products hazardous to health and environment. USA (1976), Toxic Substances Control Act (TOSCA). European Economic Community (1976), Proposal for .a council directive the Sixth modification of the council directive of 27 June 1967 on the approximation of the laws of member states relating to the classification, packaging and labelling of dangerous substances. Several of these laws and regulations are being amended and exact conditions for testing requirements etc. are being worked out. B.C.J. Zoeteman and G.J. Piet, On the nature of odours in drinking water resour-ces of the Netherlands, Sei. Total Environ., 1, 399 (1973).

1.

2.


3. F.E. Flinn, T.J. Thomas and M.D. Bishop (1974) Identification systems for se-lecting chemical classes as candidates for evaluation, EPA-650/1-74-001, Washing-ton , D.C.

4. National Academy of Sciences (1975) Assessing potential ocean pollutants, NAS, Washington, D.C.

5. V. Zitko, Comments on the status and direction of environmental research in Giam, C.S. (ed.) Biological effects of pollutants on marine organisms, D.C. Heath and Co., Lexington, Mass, (in press).

6. National Academy of Sciences (1972) Degradation of Synthetic Organic Molecules in the Biosphere, Proceedings of a conference in San Francisco, Printing and Publishing Office, National Academy of Sciences, Washington, D.C.

7. A.S.W. de Freitas, D.J. Kushner and S.U. Quadri (eds.)(1974) Proceedings of the International Conference on Transport of Persistent Chemicals in Aquatic Ecosystems, National Research Council, Ottawa.

8. R. Haque and V.H. Freed (eds.)(1975) Environmental Dynamics of Pesticides, Plenum Publ. Co., New York.

9. I.H. Suffet (ed.)(1977) Fate of Pollutants in the Air and Water Environments, Part I and II, John Wiley and Sons, New York.

10. 0. Hutzinger, G. Sundstrom, F.W. Karasek and S. Safe, The Chemistry of Some Potential Polyhalogenated Water Pollutants, in Identification and Analysis of Organic Pollutants in Water, L.H. Keith (ed.), Ann Arbor Science, Michigan, 1976.

11. Anonymous, Pollution caused by certain dangerous compounds discharged into the aquatic environment of the community, Directive of the Council of the European Communities, 18 May, 1976, Publ. Bull. L 129, p. 23.

12. R.C. Harriss (1976) Suggestion for the Development of a Hazard Evaluation Pro-cedure for Potentially Toxic Chemicals, MARC Report Nr. 3, The Monitoring and Assessment Research Centre, Chelsea College, University of London.

13. Anonymous, Law Concerning the Examination and Regulation of Manufacture etc. of Chemical Substances. Chemical Products Safety Division, Basic Industries Bureau, Ministry of International Trade and Industry, Tokyo, p. 64, 1973. S. Sasaki, Scientific aspects of the Japanese Chemical Substances Control Law, this volume.

14. F.G. Wilkes, Microcosms as Indicators of Estuarial Pollutant Stress, this volume 15. K. Maconaughey, Office of Toxic Substances, US Environmental Protection Agency,

Washington, personal communication. 16. E.T. Abrams, D. Derkics, C.V. Fong, D.K. Guinan and K.M. Slimak (1975) Identi-

fication of Organic Compounds in Effluents from Industrial Sources, EPA 560/3-75-002, Washington.

17. D.K. Guinan, R.G. Shaver and E.T. Abrams, Identification of organic compounds in effluents from industrial sources, 173rd National Meeting of the American Chemical Society, March 20-27, 1977, New Orleans, Abstracts p. 23.

18. W.B. Neely, D.R. Branson and G.E. Blau, Partition coefficient to measure bio-concentration potential of organic chemicals in Fish, Environ. Sei. Technol., 8_, 1113 (1974).

19. R.L. Metealf, J.R. Sanborn, P.Y. Lu and D. Nye, Laboratory model ecosystem studies of the degradation and fate of radiolabelled tri-, tetra- and penta-chlorobiphenyl compared with DDE, Arch. Environ. Toxicol., 3_* 151 (1975).

20. D.R. Branson, W.B. Neely and G.E. Blau, in reference 20, p. 99. 21. C.T. Chiou, V.H. Freed, D.W. Schmedding and R.L. Kohnert, Partition coefficient

and bioaccumulation of selected organic chemicals, Environ. Sei. Technol., 11, 475 (1977).

22. G.D. Veith and D.E. Konasewich (eds.) (1975) Symposium on structure-activity correlations in studies of toxicity and bioconcentration with aquatic organisms, International Joint Commission, Great Lakes Research Advisory Board, Windsor, Canada.

23. T. Fujita, J. Iwasa and C. Hansch, A new substituent, π, derived from partition coefficients, J. Amer. Chem. Soc, 86, 5175 (1964).


24. C. Hansch, A quantitative approach to biochemical structure-activity relation-ships, Accounts Chem. Res., 2j 232 (1969).

25. R. Vilceanu, Z. Szabadai, A. Chiriac and Z. Simon, Multiple structure-toxicity correlation of organic phosphorus compounds, Stud. Biophys., 34, 1 (1972).

26. C. Hansch, A. Leo, S.H. Unger, K.H. Kim, D. Nikaitani and E.J. Lien, "Aromatic" substituent constants for structure activity correlations, J. Med. Chem., 16, 1207 (1973).

27. J.L. Cohen, W. Lee and E.J. Lien, Dependence of toxicity on molecular structure: Group theory analysis, J. Pharm. Sei., 63, 1068 (1974).

28. A. Leo, C. Hansch and D. Elkins, Partition coefficients and their uses, Chem. Revs., 7JL_, 525 (1971) .

29. G. Redl, R.D. Cramer tert. and C.E. Berkoff, Quantitative drug design, Chem. Soc. Revs. , 3_' 273 (1974).

30. G. Nys and R. Rekker, Chim. Therap., 521 (1973). 31. R.F. Rekker (1977) The hydrophobic Fragmental Constant, Elsevier Publ. Co.,

Amsterdam. 32. R.G. Zepp and G.L. Baughman, Prediction of photochemical transformations of

pollutants in the aquatic environment, this volume. 33. Section II in part I of reference 9, p. 223, see also discussion on p. 478. 34. D.G. Crosby, The toxicant-wildlife complex, Pure and Applied Chemistry, 42,

233 (1975). 35. J.B. Weber, The pesticide scorecard, Environ. Sei. Technol., 11, 756 (1977). 36. P.N. Craig and J.H. Waite (1976) Analysis and trial application of correlation

methodologies for predicting toxicity of organic chemicals, Office of toxic Substances, Environmental Protection Agency, Washington, D.C., EPA-560/1-76-006; PB258119 (1976).

37. V. Zitko, unpublished results. 38. V. Zitko and O. Hutzinger, Uptake of chloro- and bromobiphenyls, hexachloro-

and hexabromobenzene by fish, Bull. Environ. Contain. Toxicol., 16, 665 (1976). 39. V. Zitko, The accumulation of polybrominated biphenyls by Fish, Bull. Environ.

Contarn. Toxicol., 17, 285 (1977). 40. G. Sundström, O. Hutzinger, F.W. Karasek and J. Michnowicz, Environmental

Chemistry of Substitutes for Polychlorinated Biphenyls -I- Composition and Properties of an Alkylchlorobiphenyl product. J. Assoc. Offie. Anal. Chem., 59_, 982 (1976).

41. M.Th.M. Tulp and O. Hutzinger, unpublished results. 42. M.Th.M. Tulp, G. Sundström, J.C. de Graaff and O. Hutzinger, Environmental

Chemistry of PCB-replacement Compounds -III- The Metabolism of 4-chloro-4'-isopropylbiphenyl and 2,5-dichloro-4'-isopropylbiphenyl in the Rat, Chemosphere, 6_, 109 (1977).

43. M.Th.M. Tulp, G.M. Tillmanns and O. Hutzinger, Environmental Chemistry of PCB-replacement Compounds -V- The Metabolism of Chloroisopropylbiphenyls in Fish, Frogs, Fungi and Bacteria, Chemosphere, 6_, 223 (1977) .

44. H.R. Sullivan, R.E. McMahon, D.G. Hoffman and S. Ridolfo, Metabolite identifi-cation by GC-MS: Species differences in the metabolic patterns of isopropyl-biphenyl. In: Mass Spectrometry in Drug Metabolism, A. Frigerio and E.L. Ghisalberti (eds.), Plenum Publ. Co., New York, 1977, p. 31.

The Environmental Chemicals Data and Information Network (ECDIN) and Related Activities of the European Communities

H. OTT*, F. GEISS** and W. G. TOWN**

* Commission of the European Communities, BrusseL·, Belgium **Commission of the European Communities, Joint Research Centre,

Ispra, Italy

ABSTRACT

The Environmental Chemicals Data and Information Network (ECDIN) is a pilot project executed within the Environmental Research Programmes of the European Communities jointly by the Commission's Joint Research Centre (Ispra Establishment) and the Indirect Action (contract research). The scope of the project is to establish a computerized file of relevant information on chemicals potentially dangerous for the environment and human health, and to design an adequate storage and retrieval software permitting a dialogue.

The system is now operational for in-house use. The files contain information for 3.500 compounds on systematic, trivial and commercial names, formulae and structure, identifiers, physical and chemical properties, analysis, production and use, distri-bution and transformation, toxicity and ecotoxicity, and regulations. Details of the file structure, examples of stored data, and the essential characteristics of the soft-ware will be given, and the facilities for substructure search will be discus-sed.

Consideration will be given to the use of data banks in regulatory action and the implementation of regulations. Further activities of the European Commission's to be discussed are the collection of mass spectra and Chromatographie data, the links of computerized spectrum libraries with substance-oriented data-banks, and the establishment of an inventory of identified water pollutants.

INTRODUCTION

The term ECDIN, which shall arise frequently in the following, stands for Environ-mental Chemicals and Information Network.

The word "Network" is rather ambitious and not justified for that which exists at present. It anticipates future developments envisaged, to be discussed briefly below.

For the time being, ECDIN is a research project, executed in collaboration between the Communities Joint Research Centre in Ispra, Italy, and a number of European Institutes and research laboratories which are under contract; these contracts are jointly funded by the Member Countries and by the Commission of the European Com-munities Environmental Research Programme, Brussels).

33

34 H. Ott, F. Geiss and W. G. Town

THE SCOPE OF ECDIN

ECDIN, once operational, will be an instrument which would enable all people engaged in environmental management and research to obtain rapidly reliable infor-mation on chemical products of environmental significance.

The basic principle of ECDIN is to store relevant information on chemicals produced in sizable quantities regardless of the form in which they are used, and their intended function; and regardless whether these chemicals are supposed to be harm-ful or not.

There is some debate as to the definition of "sizable", but the cut-off level should be somewhere around a produced quantity of 1000 kg per year.

Of course, all chemicals that are known to be highly toxic will have to be included, even if they are produced in lower quantities, and also some selected toxic natural products, as for instance, aflatoxius.

It is estimated that the number of chemicals which meet these selection criteria, will be some 20 to 30.000.

It has to be underlined that it is not intended to establish a bibliographical data bank, i.e. a document retrieval system, although data sources are usually given and reference's made to important monographs and reviews.

DATA STRUCTURE

Being aware of the enormous effort in manpower and funds involved in building up such a data bank and of the need to identify the optimum presentation of data to the customer, to design an efficient but economic storage and retrieval of data, it was decided to start with establishing a "pilot project". After about two years of planning and three years of practical work, an operational on-line system covering about 3.500 chemicals is now operational; for about 300 of them, data coverage is reasonably complete, at least for the important data categories.

This "Pilot ECDIN" is a computerized file which contains for each included product a structured "data sheet", which can be displayed entirely or in part on a video screen or on a line printer. The system is implemented on an IBM 370 computer at the Communities' Joint Research Centre at Ispra, Italy.

The data sheet is sub-divided in 10 broad categories, called macro-items. These are:

1) Identifiers 2) Physical and chemical properties 3) Toxicity and ecotoxicity 4) Chemical structure information 5) Analytical methods and analysis data 6) Supply, production and trade 7) Transport, handling, package and storaqe 8) Use and dispose 9) Dispersion and transformation in the environment 10) Regulatory data and recommendations

Each of these macro-items is subdivided in up to 20 items, as shown in Table 1.

The structure of data to be included in ECDIN, comprehensively described in an input manual, is continously reviewed in the light of the experience acquired, in order to find out the optimum structure for an operational system, and the fine structure of some items, e.g. for data on mutagenicity or effects on aquatic organisms is not yet defined.

ECDIN and Related Activities 35

TABLE 1 ECDIN data field

IDENTIFIERS

SYSTEMATIC CHEMICAL NAME (CAS PREFERRED) CHEMICAL NAMES (ENGLISH) CHEMICAL NAMES (DEUTSCH) CHEMICAL NAMES (FRANCAIS) CHEMICAL NAMES (ITALIANO) CHEMICAL NAMES (DANSK) CHEMICAL NAMES (NEDERLANDS) TRADE NAMES CHEMICAL ABSTRACTS REGISTRY NUMBER (CAS) WISWESSER LINE NOTATION (WLN) ECDIN NUMBER UNION OF EUROPEAN CUSTOMS NUMBER EEC NUMBER COUNCIL OF EUROPE NUMBER

PHYSICAL AND CHEMICAL PROPERTIES (PURE COMPOUND)

MOLECULAR WEIGHT MOLECULAR FORMULA MELTING POINT BOILING POINT DECOMPOSITION POINT DENSITY VAPOUR PRESSURE FLASH POINT COLOUR, TASTE, ODOUR SOLUBILITIES MISCELLANEOUS DATA

TOXICITY AND ECOTOXICITY

EFFECTS ON MAN EXPERIMENTAL STUDIES ON ANIMALS (TO ACCESS TOXICITY FOR HUMANS) EFFECTS ON TERRESTRIAL ANIMALS EFFECTS ON AQUATIC ORGANISMS EFFECTS ON REPRODUCTION (INCLUDING TERATOGENICITY) CARCINOGENICITY MUTAGENICITY ALLERGIC AND IMMUNOLOGICAL REACTIONS EFFECTS ON PLANTS EFFECTS ON MICRO-ORGANISMS EFFECTS ON ECOSYSTEMS ORGANOLEPTIC THRESHOLDS EFFECTS ON INANIMATE MATERIAL

CHEMICAL STRUCTURE INFORMATION

CHEMICAL STRUCTURE DIAGRAM

CHEMICAL ANALYSIS DATA

ANALYTICAL METHODS

1 .

2.

3.

4.

5.

H. Ott, F. Geiss and W. G. Town

SUPPLY, PRODUCTION AND TRADE

MANUFACTURING PROCESSES MANUFACTURING CAPACITY (BY REGION) MANUFACTURING COMPANIES PRODUCTION FOREIGN TRADE DOMESTIC SUPPLY BULK DISPLACEMENT

TRANSPORT, HANDLING, PACKING AND STORAGE

UN NUMBER IMCO CLASSIFICATION ADR CLASSIFICATION RID CLASSIFICATION IATA CLASSIFICATION UK BLUE BOOK CLASSIFICATION USA CFR 46 CLASSIFICATION WARNING SIGNS AND HAZARD CLASSIFICATION SAFETY ADVICE (S NUMBERS) COUNTER MEASURES-FIRE COUNTER MEASURES-SPILLAGE COUNTER MEASURES-MEDICAL KNOWN ACCIDENTS HAZARD INFORMATION CLEANING AND EMERCENCY SERVICES

USE AND DISPOSAL

CONSUMPTION SUMMARY USES-DIRECT USES-INDIRECT CONSUMPTION PATTERN

DISPERSION AND TRANSFORMATION IN THE ENVIRONMENT

ENVIRONMENTAL SYNTHESIS DISPERSION PATHWAYS ABIOTIC DEGRADATION BIOLOGICAL ABSORPTION,METABOLISM AND EXCRETION BIOLOGICAL RETENTION AND ACCUMULATION MONITORING OF ENVIRONMENTAL SAMPLES INTAKE BY MAN AND OTHER ORGANISMS

REGULATORY DATA AND RECOMMENDATIONS

STACK EMISSION CONTROL ENVIRONMENTAL AIR STANDARDS DRINKING WATER STANDARDS ACCEPTABLE DAILY INTAKE DIRECT FOOD ADDITIVE TOLERANCE INDIRECT FOOD ADDITIVE TOLERANCE ANIMAL FEED ADDITIVE TOLERANCES FOOD PACKAGING COMPONENT TOLERANCES PRODUCT QUALITY SPECIFICATIONS SALES RESTRICTIONS USE RESTRICTIONS OCCUPATIONAL AIR STANDARDS OTHER OCCUPATIONAL REGULATORY DATA WASTE DISPOSAL CONTROL

36

6.

7.

8.

9.

10.

ECDIN and Related Activities 37

The difficulties encountered in defining the data structure are manyfold; data on experimental toxicity studies, the basis for any risk estimation, may serve as an example. For storing data in the computer, they should be cast in a format which is as rigid as possible. This is not too difficult for data like LD 50, where one can design a table giving species, application route, vehicle, and the numerical value etc... Structuring is obviously more difficult, although not impossible, for effects on specific target organs, but it is almost impossible for clinical case reports.

Moreover, toxicologists as potential users, are not usually familiar with such a structured presentation^and they feel that it does not permit a balanced overall estimation of the risks.

Under toxicologists* influence the system contains at present, so called "condensates" i.e. short monographs specially prepared for this purpose. A revised format for toxicity data is now being elaborated which is a compromise: as far as possible, data will be squeezed in a format, but there will be space provided for information given in free text.

SOFTWARE AND RETRIEVAL OF STORED INFORMATION

In order to accelerate the implementation of the project, the SIMAS information retrieval system, which was already available locally, was used: it is sufficiently sophisticated for the pilot project, yet has, however a number of shortcomings (limited field length, access by acronyms, difficult updating); therefore, the adequacy of alternative software systems for an operational system is being explored.

There are two ways of access to ECDIN: 1) Direct access by a name or by one of the other identifiers in category 1 (see Table D.Once the compound is found, the stored information relevant to one or more of the macro-item, or items can be displayed by entering the corresponding acronyms.

2) Access via the inverted file by keywords. For this purpose, a facetted thesaurus has been developed. The keywords can be linked by Boolean logics ("and","or","and not" e t c . ) . An example for this type of inquiry could be:

Which compounds, used as insecticides, and having a aromatic ring, a nitrogroup and chlorine as structural elements are present in the system ?

On entering the acronyms corresponding to the keywords and by the symbol for linking them, the system gives the number of compounds retrieved meeting these criteria, the computer displays now on demand the names of these. Once having the names, one can go back now to a name search as explained before.

There is however, the possibility to narrow down further, before displaying the names, be adding further selection criteria. Further ways of access to the system are being investigated. One of them entering the system directly from the mass spectra created by an analytical unit.

Furthermore, work is being done which would permit to enter the system via the structural formula or substructure elements. So far, this is possible with a number of selected substructure elements used as keywords.

Chemical structures can be generated starting from the Wiswesser line notation by means of the CROSSBOW system and displayed on the video terminal.

INFORMATION ON COMMERCIAL PRODUCTS

It is obvious that storing information on mixtures, commercial products, etc, raises a number of supplementary problems.

Some mixtures, usually produced and used as such, e.g. PCB, have their own complete file. In general, however, a name search for commercial mixtures lead to the prin-cipal components,and adequate cross-references ("pointers") to the other components

38 H. Ott, F. Geiss and W. G. Town

are foreseen- Those impurities which are usually present in technical grade chemicals are handled in the same way.

E.G. ACTIVITIES RELATED TO ECDIN

Substancial activities on collecting and disseminating information on environmental pollutants were started several years ago within the framework of COST Project 64b "Analysis of organic micro pollutants in water" in collaboration with a number of European Non-Member States-These were essentially:

-the edition of an inventory of actually identified water pollutants - a computerized mass spectrum library

The main load of these excercises were carried by the Water Research Centre, Medmenham, U-K. and by the Gesellschaft für Keruforschung, Karlsruhe, Federal Republic of Germany.

The initiative for relaunching this project remained with the European Communities, and a new programme has been defined to begin shortly, to be executed as a "Con-certed Action" of the European laboratories engaged in research in the field of water pollution-It is foreseen to build up in collaboration with all participating laboratories, a computerized file of actually identified water pollutants, and to continue and intensify the establishment of the mass spectrum library.

Provisions will be made for linking these different data-bases by common unambiguous identifiers for the compounds (e.g. the CAS-number) in view of incorporating them at a long range in a network.

CONCLUSIONS

The European Communities have taken various initiatives in the field of computerized handling of data on environmental chemicals. These have to be seen in the light of a Community policy for the dissemination of scientific knowledge. The European on-line information network EURONET, to be operational in 1979, will be an efficient tool to make these data bases easily accessible throughout the Community.

Occurrence, Registry, and Classification of Organic Pollutants in Water, with Development of a Master Scheme for their Analysis ARTHUR W. GARRISON, LAWRENCE H. KEITH and WALTER M. SHACKELFORD

Environmental Research Laboratory, Athens, Georgia 30605, U.S.A.

ABSTRACT

The U.S. Environmental Protection Agency maintains a list of organic pollutants found in all types of water in North America and Europe. In June 1977, this list contained about 7000 entries of almost 1300 different organic compounds. A functional-group classification system has now been developed to show the distribution of these organic compounds among 112 functional group classes and among 29 types of water.

With this list as a nucleus, a comprehensive computerized collation and data retrieval system will be developed. This system, the Distribution Register of Organic Pollutants in Water (WaterDROP), will permit continual updating and rapid dissemination of new information on organic water pollutants. WaterDROP will contain chemical and biological descriptors for each compound and information about its concentration, source, and geographical location. Several of these parameters will be used as keys to retrieve data of particular interest to analytical chemists, epidemiologists, and enforcement officials.

The wide variety and distribution of organics in water indicates the need for development of a master analytical scheme to serve as a general protocol for analytical surveys of all volatile organic compounds present above a specified level in any type of water. A

39

40 A. W. Garrison, L. H. Keith and W. M. Shackelford

scheme now being developed will incorporate modern gas chromatography-mass spectrometry-computer techniques to identify and quantify specific organic pollutants.

INTRODUCTION

Every natural and synthetic organic compound can be expected to be found in some water sample if detection limits are lowered sufficiently (1). In fact, the number of organic compounds identified in water is increasing rapidly as scores of laboratories examine effluents, surface water, and drinking water for various reasons· This mass of data needs to be continually updated, classified, and computerized to allow development of analytical methodology, to permit accurate health effects studies, and to provide an accurate data base for setting and enforcing pollution regulations.

The occurrence and distribution of organic compounds in water is discussed in this paper, and an improved functional group classification system is described. Updated tables are presented to show the distribution of these organics among functional groups and types of water. Current U.S. Environmental Protection Agency (EPA) efforts to generate more occurrence data, develop a computerized registry system, and develop a master analytical scheme for analysis of organic pollutants in water are also discussed.

DATA COMPILATION AND REGISTRY

Current Compilations

Shackelford and Keith have published an EPA report listing organic pollutants found in all types of water in North America and Europe (2). A June 1977 update of this report contained 6944 entries of 1282 different organic compounds. The report contains lists of compounds sorted alphabetically by name, by their location or reference to a published study, and by the type of water in which they were found (29 types are included). It also includes tables that summarize the frequency of occurrence of each compound, the location or reference, and the water type. This list, however, is limited to compounds identified during survey-type analyses; in using these data, one must keep in mind that results of studies that include only analyses for specific compounds (e.g., pesticide or PCB monitoring studies) are not included. Neither does this report include quantitative data. One of the main purposes of this EPA list is to allow determination of compounds that should be included in a subset of mass spectral data to be used for more efficient computer searching and identification of organics extracted from water. The first edition of the subset, containing spectra of about 1100 water pollutants, has recently been prepared on magnetic tape.

Another valuable compilation of organic compounds found in water is that maintained by the Water Research Centre of England for the Commission of European Communities (CEC) COST-Project 64b (3). For each compound identified, the listing gives concentrations (if available), type of water sampled, date and location of sample, and reporting laboratory. Compounds are listed by major chemical or use

Classification of Organic Pollutants 41

class. Literature searches and regular submissions from laboratories participating in the COST Project 6 4b, supplemented by private communications, provide additional entries to this extensive listing. Although the bulk of the compounds listed are those found in western European waters, many data from the U.S. EPA are included. Shackelford and Keith's EPA report (2) includes the CEC data as of October 1975.

Registry-WaterDROP

A computerized library of organic compounds in water will be the ultimate answer to handling the mass of data described above. Within the next few months, the EPA will award a contract for development of such a comprehensive computerized collation and data retrieval system. This system, to be called WaterDROP (Distribution Register of Organic Pollutants in Water) will include a means for continual updating and dissemination of new information on organics in water. It will include various chemical, biological, and geographical descriptors for each compound, as well as data on concentration, source, etc., and can be accessed by several modes to search for compounds of particular interest to analytical chemists, epidemiologists, enforcement officials, and others. Figure 1 shows one possible format for WaterDROP data output (4).

The data input to WaterDROP will include:

Compound name Method of identification Synonyms Indication of confirmation CAS number Concentration Wiswesser line notation Sample site location Emperical formula Water type Molecular weight Sampling date Chemical class Literature reference Structure Source/laboratory reference Biological effects WaterDROP entry date Taste and odor threshold

The development contract will call for a feasibility study of 3 months followed by a meeting of potential users of the system to determine the best data format, the necessary computer storage and terminal facilities, and the most useful ways to retrieve data from the system. In addition to developing computer systems, the contractor will be required to prepare a handbook containing WaterDROP data in loose-leaf, hard copy form for manual use. The handbook data format may also be similar to Fig. 1.

There have been plans within the CEC to computerize the COST Project 64b list of water pollutants (3), in a manner similar to plans for WaterDROP. Communication should be maintained between the CEC and the EPA in this area, and the optimum degree of overlap between the two computer libraries should be determined. It may be best for each library to contain as much data as possible from world-wide sources to allow easy access by user countries in all parts of the world. In any case, it would be desirable for both libraries to be based upon a compatible computer language to allow exchange of programs and data. Provisions should then be made to allow interested users outside the CEC and EPA to obtain data from the


Camohor

ifcr C10 H16 O; MW: 152.26: WLN: L55 A CVTJ A A B; CAS: 000076222; TSL: EX12250

Biol. Effects: ipr-rat LD50: 900 mg/kg; sk

USOS-air: TWA 2 ppm. (Ref.

Sample Date

4-75 4-75 4-75 4-75 3-72

3-72

3-72

1-73

12-73

Sample Location

Miami, FL (USA)

Cincinnati, OH (USA)

Ottumwa, IA (USA)

Seattle, WA (USA)

Interstate Paper Corp. GA (USA)

Riceboro,

Riceboro,

Unidentified Kraft Pulp Mill, GA (USA)

Springfield,

Everett, WA

Interstate Paper Corp. GA (USA)

Weyerhaeuser Paper Co OR (USA)

Weyerhaeuser Paper Co (USA)

i -mus T D Lo : 84 mg/

TSL) .

C o n e . ( u g / 1 )

0 . 5

0 . 1

0 . 1

0 . 5

90

2 0

45

4 0 0

60

W a t e r C l a s s

D r i n k i n g

D r i n k i n g

D r i n k i n g

D r i n k i n g

W a s t e S I C : 2 6 1 1

W a s t e S I C : 2 6 1 1

Was t e S I C : 2 6 1 1

W a s t e S I C : 2 6 1 1

W a s t e S I C : 2 6 1 1

' k g ; s c u - m u s

I d e n t i f i e d

MS,

MS,

MS,

MS,

MS, I R

MS, I R

MS, I R

MS

MS

By

GC *

GC *

GC *

GC *

GC, * GC,

* GC,

*

LDLo:

L i t . R e f e r e r

1

1

1

1

2

2

2

3

3

2 2 0 0 m g / k g

C o n t r a c t i c e R e f e r e n c e

1

1

1

1

1

1

1

2

2

Appendix A: List of CODEN Abbreviations.

JAWWA5: Journal of the American Water Works Association.

Appendix B: Literature References.

1. R. G. Tardiff, W. L. Budde, W. E. Coleman, J. DeMarco, R. C. Dressman, J. W. Eichelberger, W. H. Kaylor, L. H. Keith,-· R. F. Kopfler, R. D. Lingg, L. McCabe, R. G. Melton, and J. L. Mullaney, "Organic Compounds in Drinking Water: A Five City Study", JAWWA5 (In Press).

2. L. H. Keith, "Analysis of Organic Compounds in Two Kraft Mill Wastewaters", U. S. Environmental Protection Agency Report No. EPA-660/4-75-005, Washington, D. C., 1975.

3. B. F. Hrutfiord, T. S. Friberg, D. F. Wilson and J. R. Wilson, "Organic Compounds in Pulp Mill Lagoon Discharge", U.S. Environmental Protection Agency Report No. EPA-660/2-75-028, Washington, D. C., 1975.

Appendix C: Contact References.

1. Keith, L. H., U.S. Environmental Protection Agency, Environmental Research Laboratory, College Station Rd., Athens, GA, 30601, USA. Telephone: (404) 546-3187

2. Hrutfiord, B. F., College of Forest Resources, Univ. of Washington, Seattle, WA, 98195, USA. Telephone: (206) 543-1714.

Appendix D: Standard Industrial Classification (SIC) Numbers

2611 Pulp mills.

Bornane, 2-oxo (see Camphor)

2-Bornanone (see Camphor)

2-Camphanone (see Camphor)

Fig. 1 One possible format for Water-DROP data output


libraries. An agency of the World Health Organization (perhaps the International Reference Centre for Community Water Supply), would be a logical center for distribution of this data to non-CEC/EPA countries.

WaterDROP and the CEC library must be interfaced closely with the Environmental Chemicals Data and Information Network (ECDIN) (5) and other information networks for maximum effectiveness of both types of data systems.

CLASSIFICATION AND DISTRIBUTION

Classification

It is essential that the large data base on organic compounds in water be logically organized for efficient use, with special emphasis on classification of specific organics into an array useful to those studying health effects and to analytical chemists. Classification by chemical function is most logical because chemically related compounds often have similar health effects and usually are amenable to the same analytical techniques. Chemical classification into classes of an optimum size should allow toxicologists to examine the data base more efficiently and to more effectively prioritize chemical classes for health effects studies. This process, in turn, should prioritize the classes according to needs for analytical methods development. In light of these factors, the following classification system for organics in water is proposed. This system is that of an analytical chemist; it is based upon similarities in chemical function and is broken down into narrow enough sub-classes so that one or two analytical methods will suffice for specific analysis of all members of each class. Health effects considerations are of necessity secondary because so little is known in this area.

The COST Project 64b list (3) of October 1975 was classified into 22 chemical or usage groups—this was the starting point for development of the more elaborate classification system. This system is a modification of one originally published in a consultant's report to the International Reference Centre for Community Water Supply of the WHO (6). The major modification was the elimination of three "use classes": Dyes, Pigments, and Optical Brighteners; Pesticides and Herbicides; and Surfactants. Chemicals in these classes are now included under the appropriate functional group class.

Chemical Classification System for Organic Compounds in Water

A. Major Classes

All compounds are grouped into the following 24 alphabetically arrayed major classes. All classes are of functional groups except Miscellaneous Non-Volatile Compounds.


Alcohols Aldehydes Alkane Hydrocarbons Alkene Alkyne, and Terpenoid Hydrocarbons Amides Amines Amino Acids Benzenoid Hydrocarbons Carbohydrates Carboxylic Acids and Anhydrides Esters Ethers and Heterocyclic Oxygen Compounds Halogenated Aliphatic Compounds Halogenated Aromatic Compounds Ketones Nitro-Compounds Nitrogen Compounds, Miscellaneous Non-Volatile Compounds, Miscellaneous Organometallic Compounds Phenols and Naphthols Phosphorus Compounds Polynuclear Aromatic Hydrocarbons Steroids Sulfur Compounds

B. Hierarchy of Assignment

In cases of more than one functional group in a compound (usually the case), the hierarchy of assignment to major class is as follows. For example, any compound containing a halogen atom falls into Class 2 or 3 except non-volatile compounds (see "Rules" for definitions)· Any phosphorus compound that is not halogenated falls into Class 5, even if it contains sulfur, while most sulfur compounds fall into Class 6.

1. Non-Volatile Compounds, Miscellaneous 2. Halogenated Aliphatic Compounds 3. Halogenated Aromatic Compounds 4. Amino Acids 5. Phosphorus Compounds 6. Sulfur Compounds 7. Carbohydrates 8. Steroids 9. Organometallic Compounds 10. Nitrogen Compounds, Miscellaneous 11. Carboxylic Acids and Anhydrides 12. Phenols and Naphthols 13. Amines 14. Nitro-compounds 15. Ketones 16. Aldehydes 17. Alcohols 18. Esters 19. Amides 20. Ethers and Heterocyclic Oxygen Compounds 21. Polynuclear Aromatic Hydrocarbons 22. Benzenoid Hydrocarbons 23. Alkene, Alkyne, and Terpenoid Hydrocarbons 24. Alkane Hydrocarbons


C. Sub-Classes

Within most of the 24 major functional group classes, each compound is assigned to a more narrow chemical sub-class; 112 different classification options are available. These sub-classes are listed in Table 1 with examples.

D. Rules

1. Halogens - any compound containing a halogen atom except Miscellaneous Non-Volatile Compounds (Class 1) is assigned under "Halogenated Aliphatic Compounds" or "Halogenated Aromatic Compounds".

2. The various "Aromatic" sub-classes usually include all compounds of the major class concerned that contain any aromatic function, even if the major functional group is not on the ring. Benzyl alcohol is thus listed under Alcohol/ Aromatic instead of Alcohol, Aliphatic.

3. Quinones are a sub-class under Ketones.

4. Lactones are sub-classified under Esters, Miscellaneous. Anhydrides are sub-classified under Carboxylic Acids.

5. Alkylnaphthalenes and Indene Derivatives are arranged under Polynuclear Aromatic Hydrocarbons. Indans and Tetralins, Biphenyls and Polyphenyls, and all hydrocarbon-substituted benzenes are classified under Benzenoid Hydrocarbons.

6. "Aliphatics" usually include cyclic aliphatics. "Substituted aliphatics" usually include unsaturated aliphatics, except for halogenated aliphatics.

7. "Non-Volatile Compounds, Miscellaneous", includes chlorophyll, enzymes, fulvic acids, humic acids, complex nitrogen bases, tannic acids, vitamins, xanthophylls, polymers, most of the dyes and optical brighteners, and other compounds that are not readily gas chromatographable.

8. Silicon compounds are listed under "Organometallic Compounds" (none are currently listed).

9. The order of hierarchy of functional groups for assignment to sub-classes within a major class is the same as that for assignment to major classes.

Distribution

A logical use of such a chemical classification system as developed above is in the organization and display of data on the distribution of organic compounds in water. In Table 1 the major classes and sub-classes are given along with the number of times members of each class have been found in various types of water, and the total times found. The distribution data in Table 1 come from the June 1977

TABLE 1

Distribution Of Organic Compounds In Water By Chemical Class And Water Type

CHEMICAL CLASS

Subclass -

example

Number of Compounds Found

Number

Found per of Times

Water Typ e*

TOTAL TIMES FOUND

CHEMICAL CLASS

Subclass -

example


Industrial Effluent

Municipal

Effluent

Surface Water

Ground Water

Finished Drinking Water

TOTAL TIMES FOUND

CHEMICAL CLASS

Subclass -

example


Industrial Effluent

Untreated

Treated

Surface Water

Ground Water


TOTAL TIMES FOUND

1. NON-VOLATILE COMPOUNDS, MISCELLANEOUS -

guanosine

26

14

11

2

3

0

1

31

2. HALOGENATED ALIPHATIC COMPOUNDS

167

264

16

84 549

29

549

1491

a. Aliphatic bromides, fluorides, and/or iodides -

dihromoethane

8

5

0

1

10

1

32

49

b. Aliphatic chlorides - diohlovο

ethane

31

76

9

40 236

7

196

564

c. Aliphatic chlorides w/bromine, iodine, and/or

196

564

fluorine - dihvomodiohlovο

ethane

21

12

2

14

57

2

139

226

d. Aliphatic halides, unsaturated -

hvomotviohlovo-

ethylene

28

57

2

12 175

4

59

309

e. Aliphatic halides, substituted - ethyl

ohlovo-

aeetate

27

21

0

7

5

0

26

59

f. Chlorinated cyclohexanes -

0.-BHC

5

1

1

3

8

2

8

23

g. Chlorinated polycyclic aliphatics and deriva-

tives -

dieldrin

25

75

1

0

35

9

40

159

h. Ethers, chlorinated aliphatic - 1^

2-bis(chloro-

ethoxy)

ethane

16

10

0

0

19

0

26

55

i. Miscellaneous - phosphate^

tris (ohloroethy I)

6

8

1

7

4

4

23

47

3. HALOGENATED AR0i4ATIC COMPOUNDS

183

128

16

89 241

29

194

697

a. Brominated, iodinated, and/or fluorinated

benzenes and alkylbenzenes -

dihromohenzene

4

0

0

0

8

0

8

16

>

ο

Ι-

Α

0) ο

ι-Ι

TABLE 1 Distribution Of Organic Compounds In Water By Chemical Class And Water Type (Continued)

b.

c.

d.

e.

f.

g.

h.

i.

j.

k.

1.

m.

n.

CHEMICAL CLASS

Subclass -

ex

ampl

e

Chlorinated benzenes and alkylbenzenes -

tetr

aehl

oroe

thyl

styr

ene

Halogenated aromatic amines -

2-

chlo

voan

iline

Halogenated aromatic ethers -

pe

ntao

hlor

ophe

nyl

met

hyl

ethe

r Halogenated benzoic acids -

5-

chlo

rosa

licyl

io

acid

Halogenated nitro-aromatics -

I>

Z-ni

tvoc

hlov

o-he

nzen

e Halogenated phenols and naphthols -

m

ethy

ltetr

a-ch

lovo

phen

oI

Halogenated heterocyclic aromatics -

oh

loro

-m

ethy

lqui

nolin

e DDT derivatives -

4,

4'-D

DE

Halogenated aromatic organophosphorus compounds -

ronn

e I

Halogenated atrazine derivatives -

at

vazi

ne

Halogenated phenyl- and phenoxyalkanoic acids and

esters -

2 3

4-D

3but

yl

este

v Polychlorinated biphenyls (PCB's—e.g.,

Arochlors) -

he

ptao

hlor

obip

heny

l Miscellaneous -

ch

lovo

phen

ylet

hyl

sulfo

ne

4. AMINO ACIDS -

le

ucin

e

° 0 TJ

U 3

<D 0

£ £

0 0

35

10

13

l

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9

29 8

6

2 5

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13 M-l

H W

40

15

7

14

5

27

1

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0 7

2

4

6

1 °

Number

Found per

Municipal

Effluent

untreated 7

0

1

1

0

4

0

1

0

0

1

1

0

17

Treated 34 2

6

6

0

18

5

0

0

0

5

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1

of Times

Water Type*

Surface Water 91 8

23

1

7

22

2

23

2

13

18

10

; 13

0

Ground Water 17 0

0

0

0

0

0 8

1

1

2

0 o

0


94 0

6

0

6

23

3

10

0

18

2

10

14

0

TOTAL TIMES FOUND 283

25

43

22

18

94

11

42

3

39

30

35

36

18


CHEMICAL CLASS

Subclass -

ex

ampl

e

5. PHOSPHORUS COMPOUNDS

a. Compounds with P-S bonds -

m

alat

hion

b. Phosphates and Phosphites -

tr

ibut

ylph

osph

ate

c. Miscellaneous

6. SULFUR COMPOUNDS

a. Aromatic (with S on the ring) -

th

ioph

enol

b. Compounds w/S-0 bonds -

dim

ethy

l su

lfoxi

de

c. Compounds w/S-N bonds -

N

-eth

yl-2

-tol

uene

sul-

fona

mid

e d. Heterocyclic -

di

thia

ne

e. Mercaptans, aliphatic -

t-b

utyl

m

eroa

ptan

f. Sulfides, Disulfides, and Trisulfides, aliphatic-

'

die

thyl

dis

ulf

ide

g. Miscellaneous -

eth

yl

isot

hioo

yana

te

7. CARBOHYDRATES -

43

6-d

i-o

-isopropyli

dene-L

-sorbo-

fura

no

se

8. STEROIDS -

eo

pros

tano

l

ß o

O T3 !

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<u o

kg ok

« e

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12 8

4

0

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3 ■-*

Ό 4H

H W

29

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7

1 o

63 9 5

3

! 23

i

0

19 4

2

3

Number

Found per

Municipal

Effluent

-P

ω

-P

c 3

2

1

0 3

1

0

0 2

0

0 0

12 5

CD

-P

id

<D

U

EH 1 1

0

0

9 3

0

0 4

0

0 2

0

2

of Times

Water Type*

u CD

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ω

υ

fd

m u

~

4 3

0

35 2

4

0

20 2

7

0

1

3

u 0)

-P

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9 0

16 0

0

12 4

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Pn Q

10 0

10 0

35 0

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29

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161

15

16

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29

13

17

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3

ffi

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CHEMICAL CLASS

Subclass -

ex

ampl

e

9. ORGANOMETALLIC COMPOUNDS -

co

pper

ph

thal

ocya

nine

10. NITROGEN COMPOUNDS, MISCELLANEOUS

a.

b.

c.

d.

e.

f.

g·

h.

i.

j.

k.

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am

etry

ne

Azo compounds and Hydrazines -

az

oben

zene

Carbamates -

ca

rbar

yl

Cyanides -

ben

zyl

cyan

ide

Heterocyclic -

ca

ffein

e Indoles and Carbazoles -

3-

hydr

oxyi

ndol

e Nitrosamines -

di

met

hyln

itros

amin

e Pyridines -

di

met

hylp

yrid

ine

Quinolines -

di

met

hylq

uino

line

Xanthines and Uric Acids -

1-

met

hylx

anth

ine

Other

11. CARBOXYLIC ACIDS AND ANHYDRIDES

a.

b.

c.

Anhydrides -

pht

halic

an

hydr

ide

Aromatic, benzoic and aliphatic-substituted

benzoic types -

t-

buty

lben

zoic

ac

id

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3-hy

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zoic

ac

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Ti <+H

ß M-4

H W

0

1 56 6

1 3

8

6

9 0

15 8

0

0

158 2

16

1 5

Number

Found per

Municipal

Effluent

*Ü

(D

-P

rd

-P

Ö

D 0

24

0

0 0

0

12 1 0

0 0

11 0

57

0

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n

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125 0

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u -P

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£

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υ

rd

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CO

T 24

0

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11 0 6

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104

0

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2755

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130

5555

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54

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k.

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a.

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18

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161

79

b.

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332

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2025

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MIN

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4161

73

270

15

11

3

> :E:

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3

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2

0

10

32

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3-

propanedioI

3

3

0

0

0

0

0

3

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(incld. aniline and benzidine) -

dimethyl

aniline

18

30

1

1

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5

59

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0

11

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naphthylamine

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2

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0

1

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0

3

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3

4

1

0

0

0

0

5

14 . NITRO-COMPOUNDS

23

24

1

1

17

0

12

55

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nitromethane

4

6

0

1

2

0

5

14

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Aromatic -

nitvoethoxybenzene

19

18

1

0

15

0

7

41

15 . KETONES

55

104

7

23 149

31

120

434

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propyl

ketone

25

68

6

19 132

16

88

329

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Aliphatic, substituted - methyl

propenyl

ketone

15

12

0

2

2

9

14

39

c. Quinones -

anthraquinone

4

6

0

0

6

4

4

20

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acetophenone

11

18

1

2

9

2

14

46

16 . ALDEHYDES

26

23

0

11

10

3

64

111

CHEMICAL CLASS

Subclass -

example


Surface Water

Ground Water


I TOTAL TIMES FOUND

Industrial Effluent

Untreated

Treated

Municipal

Effluent I

Number of Times

Found per Water Type*

TABLE 1


(Continued)

a. Aliphatic -

iso-hutyvaldehyde

18

6

0

8

8

3

51

76

b.

Aromatic - 2^ 4^ 6-trimethyl

henzaldehyde

4

10

0

2

2

0

7

21

c. Miscellaneous -

avotonaldehyde

4

7

0

1

0

0

6

14

17 . ALCOHOLS

67

117

9

24

40

39

61

290

a. Aliphatic -

hutanol

25

30

4

14

12

20

48

128

b.

Aromatic - a-methylhenzyl

alcohol

12

23

0

2

1

0

2

28

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a-tevpineol

11

50

2

6

26

0

9

93

d. Miscellaneous - 3-methyloyolopentan-l^

2-diol

19

14

3

2

1

19

2

41

18 . ESTERS

90

79

21

35

585

33

103

856

a. Adipates and Azelates - his

(2-ethylhexyl)azelate

9

2

0

3

16

2

9

32

b.

Alkyl alkanoates

(incld. substituted ) -

vinyl

propionate

18

14

8

7

158

0

13

200

c. Aromatic - methyl

dimethoxyhenzoate

11

14

2

0

3

0

7

26

d. Fatty acid methyl esters - methyl

linoleate

10

8

1

8

138

1

6

162

e. Phthalates - di-n-hutyl

phthalate

20

27

0

17

232

28

56

360

f. Miscellaneous - butyl

octyl

maleate

22

14

10

0

38

2

12

76

19 . AMIDES

8

8

2

3

0

8

1

22

CHEMICAL CLASS

Subclass -

example


Surface Water

Ground Water


TOTAL TIMES FOUND

Industrial

J Effluent

Untreated

J Treated

••lunicipal

Effluent

Number of Times


TABLE 1

Distribution Of Organic Compounds In Viater By Chemical Class And Water Type

(Continued) ^

a. Aliphatic and substituted aliphatic -

caprolactam

7

6

2

3

0

8

1

20

b.

Aromatic -

acetanilide

1

2

0

0

0

0

0

2

20.

ETHERS AND HETEROCYCLIC OXYGEN COMPOUNDS

51

86

6

21

49

10

48

220

a.

x^liphatic ethers - 1

1-diethoxypropane

11

7

1

6

13

7

19

53

b.

Aliphatic, substituted and/or unsaturated

ethers -

2-ethoxyethanol

3

2

0

0

0

2

2

6

c. Aromatic ethers - ethyl

benzyl

ether

17

34

0

1

10

0

11

56

d. Heterocyclic oxygen compounds - 4-methyl-l^ 3-

dioxolane

20

43

5

14

26

1

16

105

21.

POLYNUCLEAR AROMATIC

HYDROCARBONS

51

127

26

22

94

11

85

365

a. Alkylnaphthalenes

(and naphthalene) -

1-methyInaphthalene

13

66

7

14

29

2

34

152

b.

Indenes -

3-methylindene

7

14

0

0

3

2

13

32

c. Polynuclear aromatic hydrocarbons, other -

3^

4-benzopyrene

31

47

19

8

62

7

38

181

22.

BENZENOID HYDROCARBONS

45

126

12

38

160

25

204

565

CHEMICAL CLASS

Subclass -

example


ISurface Water Ground Water

Finished

Drinking Water

TOTAL TIMES FOUND

Industrial Effluent

IUntreated

Treated

Municipal

Effluent

Number of Times


TABLE 1


(Continued)

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update of Shackelford and Keith's EPA report (2), which provides the latest information available. As mentioned earlier, however, the EPA list is heavily biased in favor of survey-type analytical findings; monitoring data are not included. For example, most of the pesticides identified in various water monitoring programs conducted throughout the world during the last two decades are not included—most of the pesticides included were found recently in survey analyses for all types of organics. However, the number of different pesticides found and their chemical classifications should be nearly correct, and the number of times found in different types of water should be representative. This is true to a lesser extent for the halogenated aliphatic compounds; for example, some of the recent EPA monitoring data for selected halogenated organics are not included.

In Table 2 (a condensed version of Table 1), selected classes and groups of sub-classes of organics of special interest or potential importance are listed with their water-type distributions. Notice the small number of nitrosamines and organometallic compounds found and the relatively large number of halogenated aliphatic hydrocarbons, phenols and naphthols, and polynuclear aromatic hydrocarbons. These extremes are caused in part by biases in analytical methodology (many organometallics are of very low volatility and not amenable to GC-MS analysis) or in analytical approach; for example, many halogenated aliphatic hydrocarbons have been found because many of the analyses conducted in the last 2 years have been biased towards their detection. The low number of nitrosamines found is probably more representative of the true situation because nitrosamines have been searched for in several cases, with mostly negative results, and volatile nitrosamines would be detected by routine GC-MS survey analyses.

Table 3 shows the number of organics found in each type of water, again from Table 1 data. These occurrence data are biased simply because some types of water, such as finished drinking water, have been analyzed much more frequently than other types. The number of different organic compounds found in each sub-type of water should be fairly representative of the different functional group classes of organics present.

In Table 4, organic compounds on EPAfs list of priority pollutants (7) ("consent decree" compounds) and those on the recent National Academy of Sciences list (8) of 22 known or suspected organic chemical carcinogens found in drinking water are listed with their distribution according to water type, using the data of Shackelford and Keith (2). Pesticides on the two lists are also listed in Table 4, but their occurrence data are omitted because the data are highly biased against pesticides. Of the 91 non-pesticides listed in Table 4, 10 have not been found in any water, and 26 have not been found in finished drinking water, according to the Shackelford-Keith data. Twenty-four have been found in all types of water between one and five times only and 34 have been found only between one and five times in finished drinking water. These distributions perhaps indicate that occurrence data should have a greater weight in selection of specific compounds for monitoring and further study.

TABLE 2 Distribution Of Special Interest Classes And Groups Of Organic Compounds In Water

CHEMICAL CLASS OR GROUP

Amines, aromatic (including halogenated) (3-c;

13-c, d, & e)*

Benzene and alkyl benzenes, halogenated (3-a & b)

Carboxylic acids, aromatic (incld. halogenated) (3-e;

11-b, c, & d)

Ethers, chlorinated aliphatic (2-h)

Hydrocarbons, benzenoid (22)

Hydrocarbons, halogenated aliphatic (2-a,b,c, & d)

Mercaptans, Thiols and Sulfides (6-a,e, & f)

Non-Volatile compounds, miscellaneous (1)

Nitro- compounds (incld. halogenated) (3-f; 14-a & b)

Nitrosamines (10-g)

Organometallic compounds (9)

Phenols and Naphthols, halogenated (3-g)

Phenols and Naphthols (unhalogenated) (12)

Phenyl- and Phenoxyalkanoic acids (incld. halogenated)

(3-1; 11-j)

Phosphorus compounds (incld. halogenated) (3-j; 5)

Phthalates (18-e)

Polynuclear aromatic hydrocarbons (21)

Resin acids (11-k)

Steroids (8)

*Classes and sub-classes from Table 1 that were

used in these groupings.

3

o fa

M-l 10

0 T

3 0 U 3

<D O

Λ

QA

£

£ 3

O

& ö 1

37

39

25

16

45

88

16

26

32 3

1

29 1

91

26 |

14

20

51

10 4

r-\ cd

■H -P

U G

-P Q)

ω 3

3 rH

Τ3 M-l

Ö M-l

H pq

54

40

35

ί 10

126

150

28

! 14

29 0

0

27

195

6

29

27

127

34 3

Number

Found

Munic I per

ipal

Effluent

0)

■P

id

0)

u -P c

D 3

7

6

0

12

13 1

11 1

0 0

4 6

7

3

0

26 2 5

TJ

<D

-P

rd

<D

u H 3

34

10 0

38

67 3

2

1

0 1

18

12

10 1

17

22 0 2

of Times

Water Type

u Q) -P

rd

ω

υ

rd

M-l u 3 w 33

99 5

19

160

478

11 3

24 6

0

20

38

19 9

232

94 5 3

u 0)

-P

id

TJ

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25

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0

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5

10

28

11 1 0

u 0)

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<D C

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5

102

6

26

204

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565

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360

365

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TABLE 3 Number of Organic Compounds Found in Each Type of Water

INDUSTRIAL EFFLUENTS — 1804 entries** (Includes raw and treated)

Acrylamide (4)* Paper (169) Coal Gasification (27) Pesticide (7) Coking Works (14) Resin (3) Latex (70) Textile (70) Nylon (18) Chemicals (unspecified) (729) Oil Refining (56)

MUNICIPAL EFFLUENTS - Untreated ~ 260 entries

Domestic sewage (51) Laboratory sewage (3) Municipal incinerator aqueous wastes (12) Sewage treatment plant influent (20) Raw sewage (unspecified) (142)

MUNICIPAL EFFLUENTS - Treated — 520 entries

Chlorinated domestic sewage (9) Chlorinated sewage treatment plants (55) Sewage treatment plants (unspecified) (244)

SURFACE WATER — 2253 entries

Creeks (7) Lakes (79) Land runoff (8) Rivers (422) Sea (21) Raw water (unspecified) (146)

(for drinking water treatment)

GROUND WATER — 364 entries

Landfill leachate (62) Wells (125) Ground water (unspecified) (54)

FINISHED DRINKING WATER — 1743 entries (Includes some unchlorinated)

TOTAL ENTRIES 69 44

* Numbers of different compounds found in each water type.

** Total findings for all compounds.

TABLE 4 Distribution of Compounds on EPA Priority Pollutant List and National Academy

of Sciences List of Carcinogens

COMPOUND

acenaphthene

acenaphthylene

acrolein

acrylonitrile

faldrin

anthracene

benzene

benzidene

benzo(a)anthracene

benzo(k)fluoranthane

3,4-benzofluoranthene

benzo(ghi)perylene

benzo(a)pyrene

fBHC-alpha

tBHC-beta

tBHC-gamma(lindane)

fBHC-delta

bromoform

4-bromophenyl phenyl ether

butyl benzyl phthalate

1

carbon tetrachloride

fchlordane

chlorobenzene

>1

-P

•H 10

U -P

0 G

-H fd

U -P

P. 3

r-\

< Ή

Pi 0

*

*

*

*

*

* *

*

*

*

*

*

*

*

*

*

*

*

*

*

* *

*

to c

CD

cy> o

•H

υ

fd

u

CO

*

*

* *

*

*

* *

Number of Times

Found per Water

fd

•H -P

H ß

-P CÜ

ω 3

3 rH

ß MH

H W

7

3

0

4

2 5

0

0

1

1

1

5

3

0

0 8

3

Municipal

Effluent

T3

<D

-P

fd ω

■P c

D

0 0

0

0

0

1

1

1

4

2 3

3

0

0

1

0

0

<D

-P

fd

<D

U

EH Γ

0

0

0

0 5

0

0

2

1

1

1

1

0 1

0

3

u 0)

-P

fd o

fd

m u en 1

0

0

1

4

40 1

2

4

5

4

4

8

2 5

14

11

Type

u 0)

-P

fd

C o 1

1

0

0

1

1

0

0

1

0

1

0

0

0

5

0

4

u CD

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fd

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C

A -H

CO ,*

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24 0

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23 3

5

33

15

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76 2 3

17

12

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36

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52

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IH. o

o


of Sciences List of Carcinogens (Continued)

COMPOUND

£-chloro-m-cresol

chlorodibromomethane

chloroethane

bis- (2-chloroethoxy)methane

bis-(2-chloroethyl)ether

2-chloroethyl vinyl ether

chloroform

bis-(2-chloroisopropyl)ether

bis-(chloromethyl)ether

2-chloronaphthalene

2-chlorophenol

4-chlorophenyl phenyl ether

chrysene

t4,4f-DDT

t4,4'-DDE

t4,4'-DDD

dibenzo(a,h)anthracene

di-n-butyl phthalate

1,2-dichlorobenzene

1,3-dichlorobenzene

1,4-dichlorobenzene

3,3'-dichlorobenzidine

dichlorodifluoromethane

1 1 >1

•H CO

U -P

O G

•H fd

U -P

PM 3

< H

PM

O

W CM * ~~

* *

* *

*

* *

* *

*

*

* *

*

*

*

*

*

*

*

* *

03

ω

0 Ö

•H

Ü u fd

u

*C

S3

*

*

*

*

Number of Times

Found per Water

*->

id

•H -P

n c

-P Φ

W 3

0 4

2

2 2

3

16 2

0

0

0

0 0

1

1 5

1 5

0

0

Municipal

Effluent

Τ3 <D

-P

id

<D u -P

c

D 0 0

0

0 0

0 1

0

0

0 0

0

0

0

0 0

0 0

0

0

<D

-P

fd

<D u EH

~ö~ 5

2 0 0

0

11 0

0

0 1

0

0

0

0

1

0

2

0

0

u 0)

-P

fd

υ

rd

M-l U

= T=

7 1

0

2

0

78 9 0

0

0

0 3

1

3

5

4 5

0

0

Type

u 0)

-P

fd

ö

0 M

U

ΗΓ^

1

1

0 0

0 0

0 0

0 0

0

0

0

4

1

2 2

0

0

U

-P

fd

ω ö

•Η C

C ·Η

•Η U

ΡΜ Q

0

49 7

0 6

0

56 8

2

0 3

0

0

0

7

11 9

13 0 0

α

Ϊ3

D ο

CM

Η

ΕΜ

ΕΗ

Ο

ΕΗ 0

66

13 2

10 3

162

19 2 0

4

0 3

2

15

23

16

27 0

C

TABLE 4

Distribution of Compounds on EPA Priority Pollutant List and National Academy


COMPOUND

dichlorobromomethane

1 , 2-rliohlnroethane

1,1-dichloroethane

1,1-dichloroethylene

1,2-trans-dichloroethylene (listed as "1,2-

dichloroethylene")

2,4-dichlorophenol

1,2-dichloropropane

1,2-dichloropropylene (listed as "dichloro-

propylene")

tdieldrin

diethyl phthalate

2,4-dimethylphenol

dimethyl phthalate

4,6-dinitro-o-cresol

2,4-dinitrophenol

2,4-dinitrotoluene

2,6-dinitrotoluene

di-n-octyl phthalate

1,2-diphenylhydrazine (listed as "diphenylhy-

drazine")

tendosulfan-alpha

tendosulfan-beta

fendosulfan sulfate

| >

1

-r1

-H CO

U -P

O

G

-H id

SH -P

C4 3

·-·

< Ή

0< O

*

* *

*

*

* *

*

* *

*

*

*

*

* *

*

* * * *

CO

CD

o

■H

O

U

fd o

CO < *

*

·-·

fd

•H -P

+J <D

co 3

3 rH

Tf MH

Ö MH

H H

4 8

2

4

0

1

1

0

5

1

7

2

0 2 5

2

0

Number of Times

Found per Water

Municipal

Effluent

CD

-P

fd <u

1 -p

D 0 0

0

0

0

0

0

0

1

0

0

0

0

0 0

0

0

1

CD

fd

CD

E-i 5

0 0

0 0 1 2 0 3 0 0 0 0 0 0

0 0

n CD

f 1

CD

o

fd

m n

3

20

34 1

2

3

1

7

1

10 0

1

0

0 1 2

1

0

Type

ω

fd o u o 1 0 1

1

0

0

0

0

9 4

0

0 0

0 0

0

0

0)

fd

0) fi

Λ·Η

CO .*

•H Ö

C-H

•H M

fo Q

53

11 7

7

5

1

1

2

12 5

4

0

0 0 i

1

0

2

H EH in" <

EH O En 83

53

11

14

8

4

11

3

40

10

12 2

0 3 8

3

2

>

33

O ri

H·

GO

O 3

ffi

s pi o o



COMPOUND

tendrin

endrin aldehyde

ethylbenzene

ethylene thiourea (ETU)

bis-(2-ethylhexyl)phthalate

fluoranthene

fluorene

theptachlor

theptachlor epoxide

hexachlorobenzene

hexachlorobutadiene

hexachlorocyclopentadiene

hexachloroethane

indeno(1,2,3-cd)-pyrene

isophorone

tkepone

methyl bromide

methyl chloride

methylene chloride

naphthalene

nitrobenzene

2-nitrophenol

4-nitrophenol

N-nitrosodimethylamine

>1

-P

-H 05

U -P

0 C

-H id

U -P

CM 3

*-*

< Ή

\ß*

0 w

a* *

* *

*

* *

*

*

* *

*

*

*

*

*

* *

* *

*

* *

CO

CD

•H

Ü H id

u

w < a *l

*

* *

•

iH

rd

•H -P

U C

-P d)

CO 3

3 H

Ό M-l

Ö M-l

H W

0

5

0 4

2 6

7

0 7

1

1

4

0 3

22

20 0

4 3

0

Number

Found per

Municipal

Effluent

T>

ω

-P

<d

ω

n

-P

c

D 0 1

0

0 3

0

0

0

0 0 3

0

0

0 5 3

0

0

0

0

τ*

0)

-P

(d

<D

M

H 0 3

0 0

1

0

1

0 0

1

1

0

0

2

13 6

0

0

0

0

of Times

Water

u -P

(d

£ ω

υ

(d

m M 3

CO

0 8

0

63 5

3

4

2

0 2

3

0

0 3

30

13 2

2

0 2

Type

u -P

rd

£

T>

fi

3 o

u o 0 2

0

2

1

0

0

0

0

0

0

0

0

0

1

1

0

0

0

0

u CD

-P

(d

Ό tr»

CD C

Ä-H

to ,*

-H Ö

G-H

•H H

ΓΧ4 Q

0

12 0 9

4

1

8

0

0

9 2

4

3

9

24

15 2

0

1

0

S5

P

O

P4

cn

H §

H H 3 H o H 0

31 0

78

16

10

20 2

7

13

10 8

3

17

95

58 4

6

4

2



ON

N5

COMPOUND

• EPA Priority Pollutants

NAS Carcinogen

Number of Times

Found per Water Type

Industrial Effluent

Municipal

Effluent

Untreated

Treated

Surface Water

Ground Water


TOTAL TIMES FOUND

ο

Si

Η·

CO §

N-nitrosodiphenylamine

*

^

0

0

0

2

0

0

2

N-nitrosodi-n-propylamine

0

0

0

2

0

0

2

tPCB-1242 (Arochlor 1242)


*

tPCB-1221 (Arochlor 1221

*


*


*


*

*


*

pentachloronitrobenzene (PCNB)

•k

0

0

0

0

0

0

0

pentachlorophenol

*

9

2

1

1

0

1

14

phenanthrene

*

6

0

0

4

0

3

13

phenol

*

27

2

2

4

3

5

43

pyrene

*

2

3

1

7

1

2

16

t2, 3 , 7 , 8-tetrachlorodibenzo-£-dioxin (TCDD)

*

1,1,2, 2-tetrachloroethane

*

3

1

1

5

0

3

13

tetrachloroethylene

*

8

1

6

6

0

34

55

toluene

*

10

2

7

57

2

35

113

ttoxaphene

*

1,2, 4-trichlorobenzene

*

5

3

4

5

1

8

26

1,1,1-trichloroethane

*

0

1

3

25

0

5

34

1,1, 2-trichloroethane

*

3

0

0

1

1

5

10

trichloroethylene

tPesticides omitted.

*

*

5

1

4

70

0

10

90

trichlorofluoromethane

Available occurrence

*

0

1

2

4

0

8

15

2,4, 6-trichlorophenol

data are highly biased

*

0

0

0

0

0

3

3

vinyl chloride

against pesticides.

*

*

4

0

0

1

0

7

12


CURRENT DATA GENERATION

The EPA has several major extramural programs in progress that are designed to increase the data bank on occurrence and distribution of organic compounds in water. Most of this work involves "survey" analysis, comprehensive analysis for all the organic compounds in a sample capable of being isolated, separated, and detected by state of the art GC-MS-computer techniques. These include highly volatile, purgeable compounds; less volatile non-purgeables; and in a few cases volatile, but polar, water soluble compounds.

The main program (9), directed by the EPA's Effluent Guidelines Division, involves a survey of about 400 effluents sampled before and after treatment from 22 categories of industries to determine the levels of the so-called "consent decree" pollutants. These pollutants, now more commonly called "priority pollutants", include 114 organic compounds (listed in Table 4) that have adverse or potentially adverse health effects. This program, involving several contractors, has been active for about a year and will terminate in January 1979. Although the analytical protocol to support this program was designed to analyze the effluents specifically for the 114 organics of concern, spectral data for other organics of relatively high concentration are to be collected and stored for later identification.

The EPA is beginning to survey other types of water, such as municipal effluents and drinking water, for these priority pollutants. One survey (10), just started by Stanford Research Institute, involves the analysis of raw and finished drinking water from 400 cities for 100 pollutants selected from the priority pollutant and recent National Academy of Sciences list (Table 4). Here again, spectral data for other organics will be collected and stored for later identification.

An EPA contract with Battelle Columbus Laboratories calls for the analysis of 150 industrial effluents for all volatile organics (11). In addition, there are several projects in progress to analyze organics in effluents from energy-related industries. For example, one conducted by Research Triangle Institute and Gulf South Research Institute involves survey analysis of elements and organic compounds in aqueous and solid wastes from several energy-related industries: coal gasification and liquefaction, coal-fired power production, oil shale processing, coal mining, geothermal development, and petroleum refining. A draft of the first year's report lists elements and organics identified and quantified in 54 samples from coal gasification and oil shale processing (12). A recently completed contract with Midwest Research Institute resulted in the identification and quantification of 118 organics, including 47 organophosphorus compounds, in the effluents from 5 selected organophosphorus pesticide manufacturing plants (13).

Finally the EPA has contracted with the University of Illinois to analyze 200 samples of surface waters from industrial areas of the United States to determine the occurrence and identity of organic pollutants that may have originated from industrial discharges (14).

In all of these organic surveys, efforts are of course made to carefully catalog identities and concentrations of specific


organics, as well as their mass spectra. In addition, unidentified compounds are usually logged and their spectra stored so that recurring compounds can be spotted for more intensive spectral analysis and interpretation. A contract recently awarded by the EPA includes the development of a computer program to match the mass spectra of unknown compounds with unknowns that have been previously entered, thus allowing prioritization of frequently occurring unknown compounds for further analysis (15).

PROBLEMS AND FUTURE PROGRAMS IN ANALYSIS

Serious deficiencies exist in several areas concerned with generation of data on organics in water. Some of these are being addressed in current or planned programs within the EPA.

Master Analytical Scheme

Although much effort has been expended by several groups over the last few years on development of methodology for identification or for quantification of selected volatile (gas chromatographable) organic compounds in water, many gaps still exist. Extraction and separation conditions have not been optimized so that the recovery of any identified compound is known without further experimentation and internal reference standards have not been selected for groups or classes of compounds. Although specific techniques are available for analyzing many volatile organics in water, the methods are not comprehensive enough to cover the wide range of functional group classes of potential environmental importance; this includes virtually all volatile organic compounds.

Several protocols have been written for special situations, but there is a need to draw all the common elements together into a general scheme, incorporating the best of the latest advances in sample preparation and quantification that have been generated piecemeal. The EPA will award a contract in the last part of 1977 that will result in an analytical scheme that, when coupled with modern GC-MS-computer identification techniques, will serve as a general protocol for analytical surveys in which all volatile organic compounds above a specified level are searched for and quantified. The scheme should be a flexible guide for future work in the analysis of volatile organics in water of all types and be specific enough for ordinary situations, but adjustable by experienced personnel to fit unusually complex samples.

The work done by the contractor shall result in a master analytical scheme that can be applied as follows:

The user will be directed how to sample any industrial or municipal effluent, surface water, or drinking water in such a way as to obtain sufficient artifact-free sample for qualitative and quantitative analysis of any organic compound that will pass through a gas Chromatograph (GC) or can be derivitized to pass through a GC. Lower detection limits will depend upon compounds being analyzed for and the type of water being analyzed and will be specified in the scheme. Expected detection limits are, generally: drinking water, 0.1 yg/1; surface water, 1 yg/1; and effluents, 10 yg/1. Samples may be


grab samples or collected on accumulators. The user will be told what internal standards to add, and how and when to add them for all classes of organics in which he is interested· He will be guided in handling and preserving the liquid sample or accumulated organics. Extraction, concentration, and clean-up techniques will then be applied as specified by the scheme to allow maximum separation and recovery of the compound or class of compounds of interest to the user, or to allow survey analysis of all organics to which the total scheme applies.

The user will then be directed to use specified gas Chromatographie columns and conditions for separation and detection of all compounds of interest, first applying prescribed derivatization techniques, if necessary. Once compounds are detected, the user will identify them by previously established GC-MS-computer techniques. He will then be able to quantify all identified compounds with a computer program developed under this contract. The user will input raw data, consisting of retention times of sample components and markers, intensities of MS signals for sample components and markers, and identity and chemical class of sample components. The computer program will include stored MS detector response factors and recovery factors for marker and model compounds for each class of compound within each volatility fraction for each type of water sample. These stored data will be used to calculate, from the raw data, the estimated concentration of each compound of interest within a calculated uncertainty range. The user will apply quality control procedures, as specified by the scheme, during application of the total scheme.

The user will not be bound to use all steps of the scheme, but will be able to select segments appropriate for the analysis of a single selected compound or one class of compound. The scheme will also be flexible enough for the user to adapt it to rapid screening for specific organics with less quantitative accuracy than expected for detailed application, eliminating timeconsuming techniques designed for unusual functional group classes.

The emphasis of the contractor's experimental work will be on sampling, sample preparation, and quantification. The total scheme will be oriented towards identification and quantification by GC-MS using internal standards. Development of guidelines for the use of marker compounds (internal standards) will be part of the contract, as will recovery studies using model systems involving widely varying members of each major functional-group class of organic compound. Extraction techniques shall be optimized for each class of compound and each type of water sample, considering the desired level of detection. There will be considerable emphasis on accumulator columns; for example, the optimization of accumulator variables for intermediately volatile compounds needs to be coordinated into an applicable package. The contractor will be expected to determine optimum GC columns (including appropriate applications of capillary and packed columns) and GC conditions for each class or appropriate category of organic compound. It will be necessary for him to use state-of-the-art GC-MS-computer systems for detection and identification of compounds used in protocol

66 A. W. Garrison, L. H. Keith and W. M. Shaekelford

development, and to develop techniques for quantification, using internal standards with the MS as a detector.

Finally/ the contractor shall prove the protocol by showing its application to a variety of chemical compounds and water types. Spiked and unspiked samples of surface water, drinking water, and industrial and municipal effluents will be specified. Effluents from energy-related industries will be included. Model compounds used in recovery studies and compounds for spiking will be based on those on EPAfs priority pollutant list and others of interest in the water supply program.

The contract will require 15 man-years of work over a 20 month period. The published master scheme is expected to be available in about 2 years.

A Systematic Survey of Organic Compounds in The Environment

There is a surprising degree of ignorance regarding the identities and distribution of volatile organics in water. A systematic nationwide monitoring effort for individual organics is sorely needed in the united States This study would develop information to determine which organics are most abundant and to indicate their distribution, thus allowing research on health effects and other environmental problems to be directed towards an accurate assessment of environmental hazards, and allowing control of these hazards. Initially the survey would be limited to volatiles, but non-volatiles could be included as sampling methodology becomes available.

Analytical Methods for Non-volatiles

Because the mass spectrum of a compound provides the most definitive information currently available concerning its identity, the interface of gas chromatography-mass spectrometry-computers has naturally provided the most powerful tool available for the final step of separation and identification. In fact, more than 90% of the compounds identified in water have been detected by modern GC-MS-computer systems. These identified compounds are volatile compounds or compounds that can be made volatile (gas chromatographable) by simple chemical derivitization processes. The problem of identifying non-volatiles remain, however. Various estimates, usually based on total organic carbon (TOC) measurements, show non-volatiles to compose up to 95% of the total organic material in water. It is often postulated that the 5-25% of volatile organics in water include most of the toxic and taste and odor causing compounds expected in the total mass of organics. There is little reason to assume a priori that the non-volatiles are less hazardous than the volatiles, however.

The non-volatile area is a great unknown; most of these compounds remain unidentified because of the lack of adequate analytical methodology. Separation by liquid chromatography (LC) seems to be the key step. Liquid Chromatographie research should concentrate on separation of compounds not amenable to GC separation—GC is the preferred method for volatile and easily derivitizable organic pollutants. Most work is needed in these areas:


a. Extraction/concentration, especially by accumulator columns.

b. Separation by LC, especially for aqueous concentrates·

c. More universal and more sensitive LC detectors.

d. On-the-fly identification, analogous to GC-MS. LC-mass spectrometry (MS) shows most promise in this area, and several means of interfacing LC with MS investigated. LC-infrared and LC-Raman combinations are also being developed.

are being spectroscopy

Methods for Analysis of Sediments and Suspended or Colloidal Solids in Water

Many types of organic compounds are adsorbed on sediments and particulate, suspended, or colloidal matter in water; much of this matter is itself organic. One is always faced with the question of whether to filter the sample before extraction and, if so, whether to attempt to analyze the separated solids for organics. This may be a rare problem in finished drinking water where suspended solids are reduced to a very low level, but it is a serious question in the analysis of source water, effluents, and surface water. Much research is needed on analytical methodology for organics adsorbed on solids.

ACKNOWLEDGMENTS

Appreciation is expressed to Frances Mullins who spent many hours classifying and collating data in the tables and to Annie Smith who typed them.

REFERENCES

(1) W. T. Donaldson, Trace Organics in Water, Environ. Sei. Techno1. 11, 348 (1977).

(2) W. M. Shackelford and L. H. Keith, "Frequency of Organic Compounds Identified in Water", U.S. Environmental Protection Agency Report No. EPA-600/4-76-062, December 1976.

(3) Water Research Centre, "A Comprehensive List of Polluting Substances Which Have Been Identified in Various Fresh Waters, Effluent Discharges, Aquatic Animals and Plants, and Bottom Sediments," Stevenage Laboratory, Elder Way, Stevenage, Hertfordshire SGI ITH, England, October 1975.

(4) L. H. Keith, Distribution register of organic pollutants in water—an essential tool for environmental assessment, Preprints, Div. of Environmental Chemistry, American Chemical Society, 16, 176 (172nd national meeting, San Francisco, Aug. 29-Sept. 3, 1976") .


(5) H· Ott, F. Geiss, and W. G. Town, The Environmental Chemicals Data and Information Network (ECDIN) and related activities of the European Communities, Proceedings of 2nd International Symposium on Aquatic Pollutants, Noordwijkerhout (Amsterdam), Sept. 26-28, 1977 (in press).

(6) A· W. Garrison, "Analysis of Organic Compounds in Water to Support Health Effects Studies", WHO, International Reference Centre for Community Water Supply, The Hague, the Netherlands, Technical Report No. 9, Dec. 1976.

(7) Consent Decree, U.S. District Court for the District of Columbia, 7 June 1976.

(8) National Academy of Sciences, "Drinking Water and Health" (a report of the safe drinking water committee, National Research Council), Part II, p. VI-345 (1977).

(9) U.S. Environmental Protection Agency, "Sampling and Analysis Procedures for Screening of Industrial Effluents for Priority Pollutants", U.S. EPA, Environmental Monitoring and Support Laboratory, Cincinnati, Ohio, 45268. (March 1977, revised April 1977).

(10) J. Cotruvo, Office of Water Supply, EPA, Washington, DC, personal communication (1977).

(11) A. Alford, EPA, Environmental Research Laboratory, Athens, GA, personal communication (1977).

(12) E. D. Pellizzari, "Identification of Components of Energy Related Wastes and Effluents", draft report to the U.S. EPA on work done under Contract No. 68-03-2368 (1977).

(13) M. Marcus, J. Spigarelli, and H. Miller, "Identification of Organic Compounds in Organophosphorus Pesticide Manufacturing Wastewaters", draft report to the U.S. EPA on work done under Contract No. 68-03-2343 (1977).

(14) V. DeCarlo, Office of Toxic Substances, EPA, Washington, DC, personal communication (1977).

(15) J. McGuire, EPA, Environmental Research Laboratory, Athens, GA, personal communication (1977).

Natural Background ofAlkanes in the Aquatic Environment

S. C. BRASSELL*, G. EGLINTON*, J. R. MAXWELL* andR. P. PHILP**

* Organic Geochemistry Unit, School of Chemistry, University of Bristol, BristolBS8 ITS, England

**Chemistry Department, The University of California, Berkeley, California94720, U.S.A.

ABSTRACT

Contemporary aquatic environments generate and receive organic compounds which are of both natural and pollutant origin. The waters and sediments contain a wide range of compounds, free and bound, as insoluble debris. For example, extractable lipids are present in sediments in amounts varying from ppm to a few per cent. The various component classes - hydrocarbons, fatty acids, alcohols, etc. - can show distributions which are characteristic of the different types of aquatic environment.

Of particular interest are the hydrocarbons, especially alkanes, which are ubi-quitous in natural environments and derive from natural and from pollutant sources. The compounds determined vary in carbon number between wide limits (typically 10 - Uo) but the individual patterns are often readily correlated with the known inputs and the associated diagenetic effects. Various parameters can be used to distinguish between these inputs and can therefore assist in the recogni-tion of the effects of crude oil spills and similar events.

INTRODUCTION

Organic compounds are ubiquitous components of aquatic environments, and are of both natural and pollutant origin, contributed from a variety of sources. This paper outlines certain aspects of the origin, detection, and determination of these organic compounds as exemplified by hydrocarbons, specifically alkanes. Brief reference is made to the original literature, but the treatment is illustra-tive rather than bibliographical. Our main aim is to examine current attempts to distinguish between natural and pollutant alkanes in aquatic environments. In some situations the 'natural' input may be anthropogenic (man-induced), for example in those lakes which have become eutrophic because of their sewage input. In such instances there will be an enhanced 'natural' -algal contribution to the sediment, and although such an input is generated by a hatural' process, it is clearly un-representative of the natural background level of biological activity. Many types of compounds are contributed, such as fatty acids, sterols, ami no acids, caroten-oids and sugars, which may provide valuable information on the inputs to, and pro-cesses occurring in, the sedimentary environment. However, the majority of work so far has been concerned with hydrocarbons. As a class, they are chemically rather inert and bioresistant in comparison with functional!sed compounds. Furthermore,

69

70 S. C. Brasseil, et dl.

low temperatures and low oxygen contents drastically retard their degradation "by aerobic bacteria. Alkanes are, therefore, well-suited to provide a limited ecolo-gical record for a particular water body and its surroundings.

Alkanes are amenable to analysis by gas chromatography (GC) and computerised gas chromatography-mass spectrometry (C-GC-MS), and are relatively stable once isola-ted, unlike many other classes of biolipid and geolipid. Alkanes do not require derivatisation prior to analysis, and there is an abundant literature on GC retention data and mass spectra. Although GC is routinely used for the detection of alkanes, C-GC-MS is particularly appropriate in situations where the n-alkanes are masked by the other components present. One disadvantage of studying the alk-anes is that in some sediments they may be present in only low concentrations and therefore the possibility of sample contamination arises during analysis. The experimental procedures used for alkane analysis have been fully documented else-where (e.g. Brooks et al. , 1976; Aldridge et_al. , 1977).

Most of the published data on the distribution of alkanes in water and sediments are semi-quantitative. One problem is that the total hydrocarbon content consists of both solvent extractable and inextractable fractions. Usually, for simplicity, only the former fraction is routinely analysed. Amounts of individual fractions or individual alkanes are usually quoted as parts per million of the dry weight of sediment or of the total organic content. The total organic content as the refer-ence basis gives an indication of the relative amount of pollutant and/or natural product alkane in the sediment, but the parts per million dry weight gives a direct appreciation of the absolute amount. If the total organic carbon is used as a reference point, then the ratio is dependent on the often variable contents of bio-logical material (e.g. the presence of a single macroorganism). The quantitation is normally effected by peak measurements (GC or C-GC-MS) and adequate internal standardisation with single hydrocarbons is desirable. Other ratios, for example n~^17 .Xs-· n_("31' a r e a l s o useful in distinguishing inputs and in providing informa-tion on the extent of degradation, in relative terms. Highly complex mixtures of alkanes, for example the 'hump' observed in some petroleums and in biodegraded* petroleum fractions, present a special problem in quantitation, although the total signal of the unresolved complex mixture recorded as the 'hump' may be summed and the concentration of that general alkane type calculated.

The natural input of organic matter in aquatic environments includes both autoch-thonous and allochthonous contributions. The autochthonous input is comprised of the material generated within the basin of deposition whereas the allochthonous portion is transported into the sedimentary environment by water, wind, ice, etc. In addition, the alkane inputs will be of either direct or indirect biological origin, since the organic matter may be incorporated directly into the sediment after biosynthesis by an organism, or indirectly following modification either by biota, or by biological or chemical alteration during diagenesis and maturation. For example, the pristane entering a sediment from Zooplankton (Blumer and Synder, 1965) is an indirect autochthonous input, since it is ultimately derived from chlorophyll. In contrast, the pristane contributed from the terrestrial weather-ing of an ancient shale is an indirect allochthonous input since it has previously been incorporated into a sediment. Alkanes may also derive from pollutant sources such as crude oil spills, incomplete combustion of fossil fuels, and input of refined oils, gases and aerosols. A variety Of processes can result in the selec-tive removal, destruction and concentration of the alkanes in the environment and

*The biodegradation of an oil may occur both in its reservoir before recovery and/ or in the aquatic environment after spillage. The effects produced are similar, making the two processes difficult to distinguish.

Natural Background of Alkanes 71

so d e t e r m i n e t h e r e l a t i v e abundance p a t t e r n s o b s e r v e d i n t h e s e d i m e n t s .

I n t h e c o n c l u d i n g s e c t i o n , we examine t h e p r o b l e m of a s s i g n i n g an o r i g i n and p r o b -a b l e h i s t o r y t o t h e a l k a n e s found i n an a q u a t i c s e d i m e n t - which compounds ( and how much?) a r e of a d i r e c t b i o l o g i c a l o r i g i n , which a r e o f an i n d i r e c t b i o l o g i c a l o r i -g i n and which a r e of a p o l l u t a n t o r i g i n ? T h i s i n f o r m a t i o n i s d i r e c t l y d e p e n d e n t on t h e a n a l y t i c a l t e c h n i q u e s employed , b o t h q u a n t i t a t i v e and q u a l i t a t i v e . These i n f e r e n c e s a r e b e i n g made w i t h i n c r e a s i n g c o n f i d e n c e b u t i t i s e s s e n t i a l t h a t o t h e r p a r a m e t e r s and s o u r c e s o f i n f o r m a t i o n b e f u l l y u t i l i s e d . Fo r s i m p l i c i t y and b r e v i t y , t h e a c c o u n t h e r e o n l y d i s c u s s e s a l k a n e s . The a n a l y s e s of s e d i m e n t a r y o r -g a n i c m a t t e r a s a whole and o f o t h e r c l a s s e s o f compounds g r e a t l y e x t e n d t h e scope o f such e n v i r o n m e n t a l a s s e s s m e n t s .

AUTOCHTHONOUS ALKANES OF BIOLOGICAL ORIGIN

The a u t o c h t h o n o u s a l k a n e s o f d i r e c t b i o l o g i c a l o r i g i n found i n a q u a t i c e n v i r o n -ments a r e c o n t r i b u t e d by b o t h p e l a g i c and b e n t h i c o r g a n i s m s ( e . g . p h y t o p l a n k t o n , Zoop lank ton and b a c t e r i a ) and from r e e d s and o t h e r h i g h e r p l a n t s a l o n g t h e s h o r e -l i n e . T a b l e 1 l i s t s some examples o f a l k a n e s t h o u g h t t o b e c h a r a c t e r i s t i c o f p a r -t i c u l a r c l a s s e s o f b i o t a , a l o n g w i t h a few r e p r e s e n t a t i v e r e f e r e n c e s . I n some c a s e s t h e s e h y d r o c a r b o n s may b e u s e d a s b i o l o g i c a l m a r k e r s , c o n t r i b u t e d b y , and t h e r e f o r e i n d i c a t i v e of , c e r t a i n t y p e s o f o r g a n i s m . However , t o d a t e t h e a n a l y s e s c o v e r o n l y a s m a l l number of s p e c i e s and c o n c l u s i o n s m u s t , t h e r e f o r e , b e t e n t a t i v e .

The h i g h e r p l a n t i n p u t i s g e n e r a l l y assumed t o b e domina t ed by l e a f wax a l k a n e s , where t h e n - C p ~ , Cpg and C^-. a l k a n e s a r e i n much h i g h e r c o n c e n t r a t i o n t h a n t h e n -

Cpn and C^0 members ( E g l i n t o n e t a l . . 1962 and 1 9 6 3 ; C a l d i c o t t and E g l i n t o n , 1 9 7 3 ) .

T h u s , an abundance o f n - a l k a n e s i n t h e C~0 r e g i o n e x h i b i t i n g a h i g h c a r b o n p r e f e r -

ence i n d e x (CPl) i s t y p i c a l l y i n t e r p r e t e d as an i n p u t o f h i g h e r p l a n t m a t e r i a l .

The dominant a l k a n e p r o d u c e d by p h y t o p l a n k t o n i s n -C- , 7 (Oro e t a l . , 1 9 6 7 ; Blumer eit_

a l . , 1 9 7 1 ; G e l p i e t a l . , 1 9 7 0 ) . However , one o f t h e p r o b l e m s a s s o c i a t e d w i t h u s i n g n - C , „ a s an i n d i c a t o r f o r t h e p h y t o p l a n k t o n c o n t r i b u t i o n i s t h a t b i o d e g r a d a t i o n of

t h e l o w e r n - a l k a n e s i s f a s t e r t h a n t h a t o f t h e h i g h e r c a r b o n number n - a l k a n e s ( Johnson and C a l d e r , 1 9 7 3 ; C r a n w e l l , 1 9 7 5 ; Cardoso e t a l . , 1 9 7 6 ) : i t i s a l s o t h o u g h t t o b e d e g r a d e d f a s t e r t h a n t h e b r a n c h e d h y d r o c a r b o n s i n t h i s c a rbon number r e g i o n . A n o t h e r c h a r a c t e r i s t i c p a i r o f compounds i s t h e m i x t u r e o f 7~ and 8 -me thy l h e p t a d e c a n e s which a r e c o n t r i b u t e d by b l u e - g r e e n a l g a e (Han, 1 9 7 0 ) . C a p i l l a r y GC i s n e e d e d t o s e p a r a t e t h e s e a l k a n e s from t h e i s o and a n t e i s o a l k a n e s .

One i m p o r t a n t h y d r o c a r b o n which i s c o n t r i b u t e d by Z o o p l a n k t o n i s p r i s t a n e (Blumer and S y n d e r , 1 9 6 5 ) , a l m o s t c e r t a i n l y d e r i v e d from t h e c h l o r o p h y l l i n t h e p h y t o p l a n k -t o n d i e t o f t h e s e a n i m a l s . However , most p e t r o l e u m s a l s o c o n t a i n p r i s t a n e , a l t h o u g h i t i s e x p e r i m e n t a l l y p o s s i b l e t o d i s c e r n t h e o r i g i n of p r i s t a n e i n a q u a t i c s e d i -ments ( s e e b e l o w ) .

A b a c t e r i a l c o n t r i b u t i o n i s i n f e r r e d from a c o n t e n t o f n - a l k a n e s which have low CPI , i n t h e r a n g e C. s - Cp> , and o f m e t h y l - b r a n c h e d h y d r o c a r b o n s , c y c l o p r o p a n e

h y d r o c a r b o n s , a n d , p o s s i b l y , i s o p r e n o i d h y d r o c a r b o n s (Han, 1 9 7 0 ; Han and C a l v i n , 1 9 6 9 ; Yen, 1 9 7 5 ) : t h e r e a r e , h o w e v e r , c o n t a m i n a t i o n p r o b l e m s a s s o c i a t e d w i t h t h e c u l t u r e o f b a c t e r i a , so t h a t t h e i r p r e c i s e a l k a n e c o m p o s i t i o n s r ema in u n c e r t a i n .

O t h e r a l k a n e s e n c o u n t e r e d i n s e d i m e n t s i n c l u d e d i t e r p a n e s ( S i m o n e i t , 1975 and 1 9 7 7 ) , s t e r a n e s and t r i t e r p a n e s (Van D o r s s e l a e r e t a l . , 197*0· A l though t h e s e complex

72 S. C. Brassell, et al.

cyclic alkanes are widespread in the aquatic environment, it is thought that they are not direct biological products but are derived partially by environmental pro-cesses from alkenes, alcohols, fatty acids etc., contributed by algae, higher plants, bacteria and fungi, and partially as a result of oil pollution (see below).

AUTOCHTHONOUS INPUT OF ALKANES OF BIOLOGICAL AND NATURAL ORIGIN

There are three major natural sources of alkanes in addition to the autochthonous biological input, which contribute to aquatic sedimentary environments (Table 2). These allochthonous sources are Recent sediments and ancient sediments undergoing erosion, and the products of incomplete combustion in natural fires. The organic material and associated alkanes are transferred into the aquatic environment by means of rivers, wind, ice rafting, seeps and sediment flows. In the marine envir-onment this terrigenous material adds to, and sometimes obscures or completely masks, the marine autochthonous input.

Almost certainly, material of recent origin will suffer further extensive degrada-tion and weathering during transportation, resulting in the destruction or removal of more labile compounds. However, little work has been concerned with alkane cor-relations for original biological input, soils and river sediments and transported sediments. Undoubtedly, soil erosion, the wind-effected removal of dried-up lake-bed sediments and the processes of desert formation (desertification) transport weathered material into aquatic environments and hence to the sediments currently being deposited. Some of these processes are, of course, enhanced by man's acti-vity and in other cases, such as desertification, the precise relationships have yet to be established. Some data are available on alkanes in wind-transported dust (Simoneit, 1975; Simoneit et al., 1977) but there is much work to be done in this area in relation to the transport of organic matter. There is a need to establish the fluxes and their nature in different geographical areas with different types of soil, sediment, etc., and also the movement of sediments within their basins of deposition by processes such as bottom transport, slumping, etc. It is likely that the transport of sediments results in destruction of much of the labile organic matter, and, hence, in an apparent enhancement of the content of the relatively resistant alkanes.

Similar, but less extensive changes are to be expected in the organic content of ancient sediments undergoing weathering and transport prior to deposition in aqua-tic environments. The results of weathering on the hydrocarbon composition of oil seeps and shales has been examined in a few papers (e.g. Reed and Kaplan, 1977; Leythaeuser, 1973). Cycloalkanes such as steranes (i) and triterpanes (il) are known not to be readily utilised by microorganisms and will generally survive the weathering process. These two classes of compound may therefore be used to recog-nise an input from such sources but there is the problem of distinguishing such an input from a pollutant origin (see below). A further complexity is that the alkane concentrations and distributions of ancient sediments are dependent not only on the type of sediment (shale, coal, etc.) but also on its maturational history. The maturational or catagenetic changes in sedimentary organic matter are a result of the catalytic and thermal conditions that occur at depth and include all those chemical transformations that supercede diagenesis (see below). The degree of maturation depends on the depth of burial, the geothermal gradient and other fac-tors, such as the possible presence of nearby igneous intrusions. The variability of these factors and the time over which they may operate gives rise to examples of ancient sediments with different degrees of maturation, ranging from immature, such as the Messel shale (Kimble et__al., 197*0 9 to highly metamorphosed sediments con-taining graphite as their sole form of carbon. Thus, immature shales contain a wealth of alkanes with substantially unchanged biological marker skeletons; for example, the stereochemistry of the triterpanes reflects this low temperature his-


tory (Kimble et al., 197*0· However, other sediments, such as crude oil seeps, may have experienced a profound maturation sequence and in this case not only are many of the hydrocarbons structurally unrecognisable as being of direct biological origin but also their stereochemistry has in many cases been substantially altered.

Some rarer deposits, such as tar sands, could play a key role in particular envir-onments where the products of their weathering may dominate the allochthonous con-tribution and obscure other natural inputs.

Finally, the eolian input of alkanes from natural fires is unknown but it is pro-bable that the smoke generated by forest and grassland fires contains alkanes, as does tobacco smoke (Schlotzhauer et al., 1970).

INPUT OF POLLUTANT ALKANES

The three main inputs are indicated in Table 3. Thus, the major contemporary in-put of pollutant alkanes is known to result from spillages of crude oils and the use of refined oils in large quantities. In both cases much hydrocarbon material finds its way into the environment through drainage systems (Wakeham, 1976; Wakeham and Carpenter, 1976). In addition, there are the products of incomplete combustion, the result of man's use of fossil fuels and of wood, paper, etc. At present the total quantity of hydrocarbons, including alkanes, reaching the environment from pollutant sources, is generally held to be much larger than that afforded by the weathering of sedimentary formations, oil seeps, etc. Thus, it is difficult to locate unpolluted aquatic sedimentary environments, since most appear to have some degree of oil pollution, although there may be exceptions in the case of isolated mountain or other remote areas. Concern for the environment provides much of the impetus for research and monitoring studies in aquatic systems; it is essential, however, that effort be devoted to increasing the understanding of the natural situation, in so far as it is possible to clariiy this at present, because the distribution and fate of pollutant alkanes cannot be fully understood unless base-line evaluations are carried out.

In most aquatic environments, such as the off-shore sediments of the North Sea, the alkane content is a composite of contemporary biological input, palaeobiological input in the form of a contribution from the reworking of Recent sediments and the erosion of ancient sediments, plus a chronic background of alkanes from crude and refined petroleum sources. The latter has, of course, come from input by drainage systems, notably the Rhine, and from ships, spillages, etc. As a result of wea-thering, this background of chronic pollution is generally low but recognisable in sediments over a wide area. Extensive studies of sediments can help discern the contributions of the different components but the effects of acute pollution can be more easily seen, especially where surveys are conducted on a geographical basis in relation to a known spill and where analyses are available from cores taken at various sites. However, remobilisation of sediments by current action and slumping may result in non-sequential deposition and the best hope for an historical evalua-tion lies in core sampling at carefully selected sites, for example, some of the deep sediment traps in the North Sea. Individual oil spills require fingerprinting by GC and GC-MS techniques, so that the eventual input of weathered hydrocarbon can be related to the alkane content of the underlying sediments.

Table 3 gives a very brief summary of the situation with respect to the alkanes contributed as crude oils and refined oils. The main point is that the alkanes which are readily degraded, such as the n-alkanes, the branched alkanes and a few of the cyclic alkanes, will be recognisable against the natural background, provi-ded that weathering has not proceeded to a marked extent. However, when this is the case, then the recognition of pollutant alkanes resides in detailed examination

7ά S. C. Brasseil, et al.

of the distribution and stereochemistry of the cyclic alkanes, i.e. the steranes (i) and triterpanes (il). These are particularly resistant to biodegradation (Rubinstein et al., 1977) and their distributions can be discerned in Recent sedi-ments by C-GC-MS techniques. In particular, the presence of very complex mixtures of steranes and triterpanes with the appropriate carbon numbers and stereochemical combinations is an important aspect of detection of oil input in Recent sediments (Mülheim and Ryback, 1975 and 1977; Dastillung and Albrecht, 1976). The avail-able procedures for recognition extend, however, beyond alkanes into the series of alkene and aromatic hydrocarbons, which may also furnish valuable information about the inputs to individual sediments.

ENVIRONMENTAL PROCESSES AND THEIR EFFECT ON ALKANE PATTERNS

Certain environmental processes affect the patterns of alkanes eventually entombed in the sediments. The processes taking place are transportation, deposition and diagenesis. The consequences for the alkane fraction and for other organic com-pounds are first, fractionation, second, degradation, and third, formation by transformation of other compounds. Little is presently known of these effects in aquatic environments but a cursory summary is given in Table h. The operation of these processes will depend on the aquatic environment itself (Morris and Culkin, 1975); whether the water column is oxic or anoxic, and on the varying factors of acidity, salinity and available light etc. There is an established sequence for ease of degradation by aerobic microorganisms which may be summarised as follows:-the low molecular weight hydrocarbons are in general degraded preferentially while the straight chain alkanes are degraded faster than the branched alkanes which in turn degrade faster than the cycloalkanes (Bailey et al., 1973a and b ) . The effect of weathering is a composite of the various processes listed in Table k but microbial degradation leaves a recognisable pattern of residual alkanes, depending on the extent of the degradation. There is always the possibility of confusion resulting from the various types of hydrocarbon contribution; for example, the input of refined lubricating oils which have already had the straight chain alkan-es removed during manufacture, could be confused with the results of the microbial degradation of crude oils. Most of the processes listed in Table k are relatively short-term in respect of the times during which they modify the alkane patterns.

In geological literature the term diagenesis (Larsen and Chilingar, 1967) is usual-ly defined as the process of change which occurs in sediments after their deposi-tion, and which accompanies compaction and lithification. However, in organic geo-chemistry, it is convenient to modify this concept to include all early-stage transformations of organic compounds, from the death of the organism to the onset of changes that reflect increased temperature and pressure and may be considered as maturation. Therefore, diagenesis is comprised of changes influenced by physi-cochemical conditions (e.g. pH, Eh) and changes effected by biota. The latter category may be termed biological diagenesis, and consists of the processes of biodegradation, bi©transformation and biosynthesis, since microorganisms may des-troy, alter and generate organic matter. In addition, biota play a key role not only within the sediment, and especially at the sediment/water interface both dur-ing and after deposition, but also throughout the water column. The deposited, but biologically active, sediment may therefore be considered as an extension of the food web of the overlying water column. In Recent aquatic sediments the upper layers may represent hundreds or thousands of years of sedimentation. There may be chemical processes occurring in this time span which modify the alkane pattern of the sediments by way of a contribution from functionalised components; these changes will take place at the sediment temperature which may be in the region of 0°C - 30°C. The microenvironment within the sediment is conditioned by the sedi-ment itself and other factors which will control the reactions in relation to the formation and degradation of alkanes. There is little information about these


processes, but it does appear that the original distributions of deposited alkanes are, in general, retained - at least in part - over the above time scale.

RECOGNITION OF ALKANES OF BIOLOGICAL AND NATURAL ORIGIN AND OF POLLUTANT ORIGIN

A major difficulty is the fact that natural environments are so diverse that one cannot readily make firm inferences of a general nature. The environment of lakes, rivers, estuaries, coasts, marine basins, and the open oceans, will influence to different extents and in different ways the hydrocarbons entering the bottom sedi-ments. The extent of the alteration may help or hinder the recognition of the original input. For example, in anoxic waters, the preservation of organic matter is good and the original input is mainly retained due to reduced water circulation. In contrast, in oxic environments the alteration and degradation of organic matter make it more difficult to determine the original inputs. The different inputs are, in effect, superimposed upon one another and to some extent the ability to differ-entiate between them depends on the ability to separate and identify their specific component contributions. The best analytical approach to this problem is the use of C-GC-MS, since it enables the identification of both individual components and classes of components at the low levels in which they occur in the sediments (Wardroper et al., 1977).

For example, it is possible to identify and quantitate the full homologous series of n-alkanes in a sediment extract (Brooks et al., 1976). The higher plant input can be assessed from the relative abundances of the Cp -C ? set of alkanes, bearing

in mind the effective excess of the odd over the even carbon number homologues; any smooth (low CPl) background of a crude oil input would not then interfere. In the same way, the phytoplankton input can be estimated using n-C_ · additionally,

the 7~ and 8-methyl heptadecane peak provides a further measure of this input (Brooks et al., 1977). However, biodegradation of these shorter chain alkanes can result in a low estimate. An input from the weathering of ancient sediments may be characterised by the sterane, hopane and pristane distributions and stereo-chemistries, which also provide a measure of the maturity of such an input. These parameters may also indicate a crude oil input, which is similarly reflected by a low CPI for the n-alkanes, the range of isoprenoids and various other homologues (e.g. monocyclics) present and the existence of an unresolved envelope of alkanes, although this 'hump' can be derived from sediment pollution by fossil fuel pro-ducts (Giger and Schaffher, 1977). The ratio of the 'hump' to the n-alkanes and isoprenoids may give an indication of the extent of biodegradation. The majority of the available information on alkane distributions relates to crude oil analyses about which environmental interest is centred (Mclntyre and Whittle, 1977) with a view to the possibility of detecting, recognising and quantitating the hydrocarbons introduced by major oil spillages. Hence, the following questions are often put:-

1. Is it possible to recognise the gross pattern of petroleum hydrocarbons against the background of biological and natural hydrocarbons in the sediment?

2. Do the tests operate only for a recent spill or will they also characterise an old spill?

3. Can a recent spill be distinguished against a background of chronic oil pollu-tion?

Table 5 lists some of the parameters which may be of use in answering such ques-tions on the basis of the alkane content of the sediment.

The answers to all 3 questions can be 'yes' - but only in favourable circumstances,


as much depends on the characteristics of the oil input, and the relative types and quantities of the natural inputs, both autochthonous and allochthonous. Also, at the present time, reference analytical data are often too limited. For example, the stereochemistry of pristane may well be useful in discerning the input of org-anic matter from matured sediments but sufficient stereochemical analyses of rep-resentative organisms, Recent sediments and ancient sediments are not yet avail-able.

In summary, enough is known already of the natural background of alkanes in aqua-tic environments for analyses to be of some value in assessing chronic and acute pollution by crude and refined petroleums. However, there is an urgent need to make detailed surveys of selected environments so that more precise information can be obtained on the inputs of individual alkanes and on the identities and rates of the processes which control their relative abundances. In particular, studies of largely unpolluted environments will be crucial in establishing the natural levels of individual alkanes in sediments to act as a baseline reference.

ACKNOWLEDGEMENTS

The authors thank the Natural Environment Research Council (U.K.) and the National Aeronautics and Space Administration (U.S.A.) for support, and Professor P.A. Schenck, Dr. J. de Leeuw, J.J. Boon and Dr. S.G. Wakeham for helpful discus-sions and comments.

R = H,CH(CH3)2t

CH(CH3)(CH2)nCH3

Steranes

II (n = 1 - 5)

triterpanes (hopane-type)

R

III

22R,S lToiH-homohopane

R = H,CH3,C2H5

rearranged steranes


TABLE 1 Autochthonous Input of Biological Origin

TABLE 1A - n-Alkanes

Organism Environment Dominant Carbon No(s) .

CPI Carbon No. Range 14-29

Modality Example References

photosynthetic bacteria

non-photo-syn the t i c b a c t e r i a

fungi

b lue-green algae

algae

brown algae

red algae

Zooplankton

higher plants

Aquatic (Pelagic)

Aquatic (Benthic)

Aquatic (Pelagic)

Aquatic (Pelagic)

Aquatic (Benthic)

Aquatic (Benthic)

Aquatic (Pelagic)

Terrestrial

C i r C25 Low

C1T-C20 Low

C1T & C25 Low

bimodal Han, 1970

'29

'IT

17

'15

'17

High

High

Low

Low

C18 & C2k L o W

C 2 r C29 High

15-28 unimodal Han, 1970

15-29 bimodal Han, 1970

25-29 unimodal Yen, 1975

lU-19 unimodal Blumer et_ al., 1971

15-21 unimodal Gelpi _ejt al., 1970

13-26 unimodal Youngblood e t a l . , 1973

15-24 unimodal Youngblood e t a l . , 1971

18-34 or bimodal Giger and 20-28 Schaffner,

1977 15-37 unimodal Cald icot t

o r C31

TABLE IB - Branched and Cyclic Alkanes

Organism Environment Compound Range

ton, 1973

Φ Example References

Monomethyl b a c t e r i a b lue-green algae

h igher p l a n t s

Isoprenoid b a c t e r i a

algae

Zooplankton

Cyclopropane

algae

Aquatic

Aquatic (Pelagic)

Terrestrial

Marine (Benthic) Marine (Pelagic)

Aquat i c (Pelagic)

Marine (Benthic)

iso-alkanes

methyl-heptadecanes

i s o - and an te i so -a lkanes

phytane

p r i s t a n e

p r i s t a n e

17 cyclopropane

C25-C31 ■Ί8

Yen, 1975 Han and Calvin, 1970

Cp -C ? - Eglinton e t a l . , ° J 1962, 1963

"20

'19

'19

'17

Han and Calvin, 1969a

Blumer et al., 1971

Blumer and Snyder, 1965

Youngblood et al., 1971

?

78 S. C. Brassell, et al.

TABLE 2 Allochthonous Input of Alkanes of Indirect Biological and Natural Origin

Alkane class Dominant compound

Carbon No. Range

Comment Example References

Recent Sediments1 ( i n d i r e c t b i o l o g i c a l o r i g i n 2 ) s o i l s , r i v e r s and lakes

n-alkanes

branched alkanes isoprenoid

t r i t e r p a n e s

c C C υ17* 29* 31

-18

C19'°20

C 30 ' C 31 restricted

distribution

C13 C35

18

C19'C20

high CPI, urn- or bimodal

several isomers

C (?),C (?),hopanes 17PH- and ' * 17CXH-; both only one

30' 31 C-22 isomer for C3]_

Brooks et al., 1976.

Brooks et al.,

1977.

Brooks et al., 1977

Brooks et al., 1977 Van Dorsselaer et_ al., 197U

Ancient Sediments (indirect 'natural' origin )

(i) Shales^ and coals as for Recent sediments (above), but in addition the following cyclic alkanes:-

steranes

t r i t e r p a n e s

t e t r a t e r p a n e s

C 27 ' C 28 ' C 29

C 30 ' C 31

C27 °29

C27 °32 (and above)

C27~C32

limited distribution of stereoisomers

173H-(only one C-22 isomer above C_ )

I701H-dependent on matur i ty

Mülheim and Ryback, 1975

Van Dorsselaer e t a l . , 1 9 7 ^ Ensminger e t a l . , 197^

^ Q l im i t ed Murphy et a l . , mixture 1967; Kimble e t

a l . , 197^ (ii) Oil seeps as for 'crude oils' (Table 3) except that input from oil seeps

i s , in gene ra l , h ighly weathered ( l o s s of v o l a t i l e s and n -a lkanes , branched and cyc l i c a lkanes , r e s u l t i n g in r e s i d u a l unresolved complex mix tu res , UCM)

Reed and Kaplan, 1977

Products of Incomplete Combustion

Forest fires; grassland fires. Smoke may contribute unburnt alkanes.

Footnotes

1. The results available for the alkane composition of soils and river sediments are insufficient for separate documentation. 2. For convenience in classification, ancient sediment contributions of alkanes to the Recent sediment have been ignored and hence the alkanes are classed as allochthonous, i.e. contributed by transport from their site of origin to the site of deposition. 3. These data are for immature shales, which have not experienced raised tempera-tures to any appreciable extent. Erosion of such sediments frequently contribute material to contemporary aquatic environments. Prominent examples include the well-studied Green River (e.g. Anders and Robinson, 1971) and Messel (e.g. Kimble et al., 197*0. Mature sediments would be classified with crude oils (Table 3). However,

there exists a complete range of sediments between these two extremes of maturity. k. Other triterpanes are encountered, e.g. gammacerane (Hills et al. , 196*6) and moretane (Wardroper et al., 1977), but the hopane series are the principal ones.


TABLE 3 Input of Pollutant Alkanes

Alkane class Dominant compound

Crude Oils

n-alkanes

branched alkanes

v a r i a b l e

va r i ab l e

isoprenoids C.Q,Cpn

'hump'

s t e ranes ( i ) CP7~C28

rearranged C27~C28 s te ranes (IV)

t r i t e r p a n e s ^ COQ,C^n

( I I ) ^ 3 °

Carbon no.range

Comment Example References

C10 C35

C10 C35

Low CPI, uni- or bimodal, often maximising £\2~^2.§ Tiss°t et al. ,1977 region

Multiple homologies, smooth distributions, iso- and anteiso-etc.

Tissot et al.,1977

C12 C25 Range of pseudo- Han and Calvin, homologies, C-JQ i s mixed 1969b

C0£,Co A,CQ n 6RS,10RS (Va-cj Haug and Curry, 197^

u26jU28'u30

C10 C35

C27 C29

C27 °29

C27 C35

Unresolved complex mix-ture (UCM)7 smooth enve- Walker et al. ,1973 lope,uni- or bimodal

Complex mixture of stereoisomers

Ser i e s (IV)

Mülheim and Ryback, 1975 and 1977 Rubinstein et a l . , 1977 Albaiges;Rubinste in (unpublished data)

Hopanes; 17oH-configura- Das t i l l ung and t i o n with £a . equimolar Albrecht , 1976 C-22 d ias te reoisomer p a i r s for C3^-035 ( e g . I l l )

Refined o i l s

( e . g . l u b r i c a t i n g o i l s , n - a lkanes , branched a lkanes , i sop reno ids , _ *, cyc l i c_a lKa . e s ) . _ _ _ ^ g ~ « Composition modified from t h a t of crude o i l s by dewaxing (removal — " ' of n - a l k a n e s ) , d i s t i l l a t i o n e t c . (Narrowing of carbon number r ange ) .

Products of Incomplete Combustion

( e . g . f o s s i l f u e l s , wood, paper , e t c . ) Smoke may con t r ibu te unburnt a lkanes .

Footnotes

1. The crude oil component alkanes are a selection based on their ubiquity and on relative abundance of oils and their ease of separation and/or analysis. Oils vary considerably in the relative proportions and carbon number range of each class of alkane and in the distributions of constituent homologies. For example, crude oils may contain major or minor amounts of steranes and triterpanes: however, even minute contents of these bio-resistant alkanes may become proportionately more pro-minent in weathered oils as a result of preferential removal of other compounds, and their distribution may be used as a 'fingerprinting' technique (Aldridge et al., 1977).

2. The carbon number range observed in the analysis does depend on the experimental procedures, which produce a loss of the lower homologues.

3. Other triterpanes are encountered in crude oils, e.g. ΐδα-oleanane (Smith _e_t al. , 1970), but the hopane series are the most widespread in occurrence.

none

e.g. Kuras et al. , 1976

00

o

TABLE

h

Environmental Processes and their Effect on Alkane Patterns

Process

Site

Changes

Result

Example References

Photoxidation

Evaporation

Solution

Particle

association

Microbial

degradation

Microbial

alteration

Diagenesis

(chemical)

surfaces and

upper water

column

water/air

interface

water column

water column

and sediment

water column

and sediment

water column

and sediment

sediment

loss of alkanes

loss of low MW hydrocarbons

loss of lower MW hydrocarbons

higher plant debris remains in

coarse fractions; some alkanes

adsorbed onto clay surfaces

alkane oxidation

transformation of functional-

ised compounds

transformation of functional-

ised compounds

alkane patterns

changed

reduced abundance

of < C]^ alkanes

reduced abundance

of < C20 alkanes

partial fractiona-

tion

loss of n-alkanes

> branched >

cyclic

generation of

alkanes?

generation of

alkanes?

Hansen, 19T5

Milner et al., 1977

Milner et al., 1977

Thompson and Eglinton,

in press

03

to

Bailey et al., 1973a and

o

b

Philp et al., in press

Philp et al., in press


TABLE 5 Recognition of Alkane Inputs to Recent Sediments

TABLE 5A Straight chain, Branched and Isoprenoid Alkanes

Alkane parameter Characterisation of parameter

Likely inference for input J-

n-Alkanes

present

dominant homologue(s)

range of homologues

abundance r a t i o of homologues

Branched alkanes dominant component

range of homologies ( m u l t i p l e ; i s o - , a n t e i s o - , e t c .

high r e l a t i v e abundance low r e l a t i v e abundance

C ^ and/or C ^ ; high CPI C 2T' C 29» C 31 ' h i S h C P I

C 10" C 35 ' l o w C P I> unimodal or bimodal c1T/c31 » i c1T/c31 « i

C2Q/Clh > 1

7- and 8-methyl heptadecanes

c10~c353

butions

smooth distri-

Isoprenoids

range of pseudo-homologues ^\\f^25

dominant component p r i s t a n e and/or phytane

s tereochemist ry of p r i s t a n e 6R,10S-isomer alone (Va)

6RS,10RS-isomers (Va-c)

abundance r a t i o p r i s t a n e + phytane n-C IT

» 1

little biodegradation extensive biodegradation

algae

higher plant

crude oil

a l g a l alkane major h igher p l an t alkane major weathered crude o i l ( for low CPI)

b lue-green algae

crude o i l

crude o i l b i o l o g i c a l , occas ional ly crude o i l b i o l o g i c a l , ancient s e d i -ment crude o i l

p a r t i a l l y biodegraded crude o i l £ r l u b r i c a t i n g o i l

Footnotes 1. Several interpretations can often be placed on a given parameter but only one is given here. A better appreciation is dependent on the consideration of several parameters, dovetailed to minimise ambiguity. Even so, it may not be possible to discriminate between two or more possibilities. Environmental assessments fre-quently make use of data for chemical classes other than alkanes.


TABLE 5B Cyclic Alkanes

Alkane parameter

range of homologues (e.g. multiple monocyclic)

unresolved complex mixture (UCM, fhumpf)

abundance ratios (peak heights)

sterane distributions

(I, C27-C29)

rearranged sterane

distributions (IV, ^21"^2^

hopane triterpanes

(II, ^2Y~Coc)

Characterisation of parameter

Likely inference for input1

C-^Q-C^C, smooth d i s t r i -bu t ions

crude o i l

C J Q - C ^ C , smooth envelope crude o i l or f o s s i l fuels

'hump1/n-C2l+ > 1 'hump ' /p r i s tane + phytane

> 1 absent

present, simple mixture present, complex mixture

absent present

absent

present as limited com-ponents, especially C301T3H and C311T3H and Cß-LlTaH (one C-22 isomer each)

C27""C35 Present» ΙΤθίΗ-series/one C-22 isomer

C2T-C0C present - lTotH-series ( a. 1:1 pairs of C-22 isomers)

C3 1 isomers (22RS-):III short R^/long Rrp < 1

short Rrp/long Rrp ~ 1

biodegraded crude oil

extensively biodegraded crude oil

no crude oil or ancient sediment^ ancient sediment^ crude oil

no crude oil crude oil

no crude oil or ancient sediment Recent sediment (diagene-tic origin)^ no crude oil

ancient sediment

crude oil

mixed Recent sediment and crude oil

crude oil

Footnotes

1. Several interpretations can often be placed on a given parameter but only one is given here. A better appreciation is dependent on the consideration of several parameters, dovetailed to minimise ambiguity. Even so, it may not be possible to discriminate between two or more possibilities. Environmental assessments fre-quently make use of data for chemical classes other than alkanes.

2. Ancient sediment implies immature sediment (Table 2), possibly present, as single C-22 isomers.

3. The abundances of the higher homologues (C^j 3^5 c^k with the consequence that these peaks may not be observed.

ΙΤθίΗ-series also

drop off steeply,


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These references provide access t o the appropr ia te l i t e r a t u r e and are an i l l u s t r a -t i v e c r o s s - s e c t i o n r a t h e r than a d e f i n i t i v e b ib l iog raphy .

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D.E. Anders and W.E. Robinson, Cycloalkane c o n s t i t u e n t s of the bitumen from Green River s h a l e , Geochim.Cosmochim.Acta 355 66l (1971)·

N.J .L . Ba i l ey , A.M. Jobson and M.A. Rogers, B a c t e r i a l degradat ion of crude o i l : comparison of f i e l d and experimental d a t a , Chem.Geol. 1 1 , 203 (1973a).

N.J .L . Ba i l ey , H.R. Krouse, C.R. Evans and M.A. Rogers, A l t e r a t i o n of crude o i l by waters and b a c t e r i a - evidence from geochemical and i so tope s t u d i e s , Amer. Assoc .Pe t ro l .Geo l .Bu l l . 57, 1276 (1973b).

M. Blumer and W.D. Snyder, I soprenoid hydrocarbons in Recent sediments: presence of p r i s t a n e and probably absence of phytane , Science 150, 1588 (1965).

M. Blumer, R.R.L. Gui l la rd and T. Chase, Hydrocarbons of marine phytoplankton, Mar.Biol. 8, 1Ö3 (1971).

P.W. Brooks, G. Eg l in ton , S . J . Gaskel l , D.J. McHugh, J .R. Maxwell and R.P. P h i l p , Lipids of Recent sediments Pa r t I : S t r a i g h t chain hydrocarbons and carboxy-l i c acids of some temperate and s u b - t r o p i c a l l a g o o n a l / t i d a l f l a t sediments , Chem.Geol. l 8 , 21 (1976).

P.W. Brooks, G. Eg l in ton , S .J . Gaske l l , D.J. McHugh, J .R . Maxwell and R.P. P h i l p , Lipids of Recent sediments Pa r t I I : Branched and c y c l i c alkanes and alkanoic acids of some temperate l a c u s t r i n e and s u b - t r o p i c a l l a g o o n a l / t i d a l - f l a t s e d i -ments, Chem.Geol. 20, 189 (1977).

A.B. Cald ico t t and G. Eg l in ton , Surface waxes, in Phytochemistry 3 , Inorganic Elements and Specia l Groups of Chemicals (ed. L.P. M i l l e r ) , Van Nostrand Reinhold, N.Y., 162 (1973).

J . Cardoso, P.W. Brooks, G. Eg l in ton , R. Goodfellow, J .R . Maxwell and R.P. P h i l p , Lipids of r ecen t ly -depos i t ed a l g a l mats a t Laguna Mormona, Baja Ca l i fo rn i a , in Environmental Biogeochemistry (ed. J .O . Nr iagu) , Ann Arbor Sc ience , Mich., 1U9 (1976).

P.A. Cranwell, Decomposition of aqua t ic b i o t a and sediment formation: organic compounds in d e t r i t u s r e s u l t i n g from microbia l a t t ack on the a lga Ceratium h i r u n d i n e l l a , Freshwater B io l . 6, kl (1976).

P.A. Cranwell, Organic geochemistry of Cam Loch (Sutherland) sediments , Chem.Geol. 20, 205 (1977).

M. Das t i l lung and P. Albrecht , Molecular t e s t for o i l p o l l u t i o n in surface s e d i -ments, Mar .Po l lu t .Bu l l . 7, 13 (1976).

G. Eg l in ton , R.J . Hamilton, R.A. Raphael and A.G. Gonzalez, Hydrocarbon c o n s t i t u -ents of the wax coat ings of p l an t l e a v e s : a taxonomic survey, Nature 193, 739 (1962).

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G. Eglinton and R.J . Hamilton, The d i s t r i b u t i o n of a lkanes , in Chemical P lan t Taxonomy (ed. T. Swain), Academic P r e s s , 18? (1963).

A. Ensminger, A. Van Dorsse laer , C. Spyckere l l e , P. Albrecht and G. Ourisson, Pen tacyc l ic t r i t e r p e n e s of the hopane type as ubiqui tous geochemical markers: o r ig in and s i g n i f i c a n c e , in Adv.in Org.Geochem.1973 (eds . B. T i sso t and F. Bienner ) , Edi t ions Technip, P a r i s , 2*+5 (197*0.

E. Gelpi , H. Schneider , J . Mann and J . Oro, Hydrocarbons of geochemical s i g n i f i c -ance in microscopic a l g a e , Phytochemistry 9 , 603 (1970).

W. Giger and C. Schaffner, A l i p h a t i c , o l e f i n i c and aromatic hydrocarbons in Recent sediments of a h ighly eutrophic l a k e , in Adv.in Org.Geochem.1975 (eds . R. Campos and J . Goni), ENADIMSA, Madrid, 375 (1977).

J.C-Y. Han, Chemical s tud ies of t e r r e s t r i a l and e x t r a t e r r e s t r i a l l i f e , Ph.D. Thes i s , Univers i ty of Ca l i f o rn i a , Berkeley, Ca l i fo rn i a , U.S.A. (1970).

J . Han and M. Calvin, Hydrocarbon d i s t r i b u t i o n of algae and b a c t e r i a and microbio-l o g i c a l a c t i v i t y in sediments , Proc.Nat.Acad.Sci.USA 6k 9 U36 (1969a).

J . Han and M. Calvin, Occurrence of Cp -Cpc ; i soprenoids in Bel l Creek crude o i l ,

Geochim. Cosmochim. Acta 33, 733 (1969b).

J . Han and M. Calvin , Branched alkanes from blue-green a lgae , J.Chem.Soc.Chem. Commun., l*+90 (1970).

H.P. Hansen, Photochemical degradation of petroleum hydrocarbon surface films on sea wa te r , Mar.Chem. 3 , 183 (1975).

P. Haug and D.J. Curry, Isoprenoids in a Costa Rican seep o i l , Geochim.Cosmochim. Acta 38, 601 (197*0.

I .R. H i l l s , E.V. Whitehead, D.E. Anders, J . J . Cummins and W.E. Robinson, An o p t i c a l l y - a c t i v e t r i t e r p a n e , gammacerane, in Green River5 Colorado, o i l -shale bitumen, J.Chem.Soc.Chem.Commun. 20, 752 (1966).

R.W. Johnson and J .A. Calder , Ear ly d iagenes is of f a t t y ac ids and hydrocarbons in a s a l t marsh environment, Geochim.Cosmochim.Acta 37, 1953 (1973).

B . J . Kimble, J .R. Maxwell, R.P. P h i l p , G. Eg l in ton , P . Albrech t , A. Ensminger, P. Arpino and G. Ourisson, T r i - and t e t r a t e r p e n o i d hydrocarbons in the Messel o i l s h a l e , Geochim.Cosmochim.Acta 37, I I65 (197*0.

M. Kuras, M. Ryska and J . Mostecky, Analysis of s a tu r a t ed hydrocarbons by f i e l d i on i za t i on mass spectrometry, Anal.Chem. *+8, 196 (1976).

G. Larsen and G.V. Chi l ingar ( e d s . ) , Diagenesis in sediments , E l s e v i e r , London (1967).

D. Leythaeuser , Effects of weathering on organic mat ter in s h a l e s , Geochim. Cosmochim.Acta 37, 113 (1973).

A.D. Mclntyre and K.J . Whi t t le ( e d s . ) , Petroleum Hydrocarbons in the Marine Environment, ICES, 171 , Denmark (1977).

C.W.D. Milner , M.A. Rogers and C.R. Evans, Petroleum t ransformat ions in r e s e r v o i r s , J.Geochem.Explor. 7, 101 (1977).


R.J. Morris and F. Culkin, Environmental organic chemistry of oceans , f jords and anoxic b a s i n s , in Environmental Chemistry, V o l . 1 , (Sr . Rep.G. E g l i n t o n ) , The Chemical Society S p e c i a l i s t P e r i o d i c a l Repor t s , London, 8l (1975)·

L . J . Mülheim and G. Ryback, Stereochemistry of some s t e ranes from geologica l sources , Nature 256, 301 (1975).

L . J . Mülheim and G. Ryback, I s o l a t i o n and s t r u c t u r e ana lys i s of s t e ranes from geo-l o g i c a l sources , in Adv.in Org.Geochem.1975 ( eds . R. Campos and J . Goni), ENADIMSA, Madrid, 173 (1977).

M.T.J. Murphy, A. McCormick and G. Eg l in ton , Perhydro-ß-carotene in the Green River s h a l e , Science 157, 10U0 (1967).

J . Oro, T.G. Toraabene, D.W. Nooner and E. Gelpi , A l ipha t i c hydrocarbons and f a t t y acids of some marine and freshwater microorganisms, J . B a c t e r i o l . 9 3 , l 8 l l (1967).

E.B. Overton, J . Bracken and J . L . Lase t e r , Appl ica t ions of g lass c a p i l l a r y columns t o monitor petroleum-type hydrocarbons in mature sediments , J .Chrom.Sei . 15 , 169 (1977).

R.P. P h i l p , S. Brown, M. Calvin , S. B r a s s e l l and G. Eg l in ton , Hydrocarbon and f a t t y acid d i s t r i b u t i o n s in r ecen t ly -depos i t ed a l g a l mats a t Laguna Guerrero, Baja Ca l i f o rn i a , in Environmental Biogeochemistry and Geomicrobiology (ed. W.E. Krumbein), Ann Arbor Sc ience , Michigan, in p r e s s .

W.E. Reed and I .R. Kaplan, Thechemistry of marine petroleum seeps , J.Geochem. Explor. 7 , 255 (1977).

I . Rubins te in , 0. S ieskind and P. Albrech t , Rearranged s t e r enes in a s h a l e : occurrence and simulated formation, J.Chem.Soc. Perkin T r a n s . I . , 1833 (1975).

I . Rubins te in , O.P. S t r a u s z , C. Spyckere l l e , R .J . Crawford and D.W.S. Westlake, The o r ig in of the o i l sand bitumens of Alber ta : a chemical and a microbia l s imulat ion s tudy, Geochim.Cosmochim.Acta kl9 13^1 (1977).

W.S. Schlotzhauer , I , Schmeltz and S.F. Osman, Evidence for the o r ig in of mono-alkenes in c i g a r e t t e smoke, Chem.Ind. 1*3, 1377 (1970).

B.R.T. Simoneit , Sources of organic mat te r in oceanic sediments , Ph.D. Thes i s , B r i s t o l (1975).

B.R.T. Simoneit , Di terpenoid compounds and o the r l i p i d s in Deep-Sea sediments and t h e i r geochemical s i g n i f i c a n c e , Geochim.Cosmochim.Acta Ul, U63 (1977).

B.R.T. Simoneit , R. Chester and G. Eg l in ton , Biogenic l i p i d s in p a r t i c u l a t e s from the lower atmosphere over the eas t e rn A t l a n t i c , Nature 267, 682 (1977).

G.W. Smith, D.T. Fowell and B.G. Melsom, Crys ta l s t r u c t u r e of l8otH-oleanane, Nature 228, 355 (1970).

S. Thompson and G. Eg l in ton , The f r a c t i o n a t i o n of a sediment for organic geochemi-ca l a n a l y s i s , Geochim.Cosmochim.Acta, in p r e s s .

B. T i s s o t , R. P e l e t , J . Roucache and A. Combaz, U t i l i s a t i o n des alcanes comme f o s s i l e s geochimiques i n d i c a t e u r s des environnements geologiques , in Adv. in Org.Geochem.1975 (eds . R. Campos and J . Goni) , ENADIMSA, Madrid, 117 (1977).


A. Van Dorsselaer, A. Ensminger, C. Spyckerelle, M. Dastillung, 0. Sieskind, P. Arpino, P. Albrecht, G. Ourisson, P.W. Brooks, S.J. Gaskell, B.J. Kimble, R.P. Philp, J.R. Maxwell and G. Eglinton, Degraded and extended hopane deri-vatives (Cp7 to C ,_) as ubiquitous geochemical markers, Tetrahedron Lett. ,

13^9 (197*0.

S.G. Wakeham, A comparative study of petroleum hydrocarbons in lake sediments,

Mar.Pollut.Bull. 7, 206 (1976).

S.G. Wakeham and R. Carpenter, Aliphatic hydrocarbons in sediments of Lake Washington, Lim. and Oceanography 21, 711 (1976).

J.D. Walker, R.R. Colwell and L. Petrakis, Microbial petroleum degradation: application of computerized mass spectrometry, Can.J.Microbiol. 21, 1760 (1975).

A.M.K. Wardroper, P.W. Brooks, M.J. Humberston and J.R. Maxwell, Analysis of steranes and triterpanes in geolipid extracts by automatic classification of mass spectra, Geochim. Cosmochim. Act a 1+1, U99 (1977)·

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W.H. Youngblood and M. Blumer, Alkanes and alkenes in marine benthic algae, Mar.Biol. 21, 163 (1973).

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Multidetection Approach to Analysis of Organic Pollutants in Water. Methods and Comments on Results

R. FERRAND, M. MAZZA and P. PAYEN

CERCHAR, B.P. No 2, 60550 Verneuil-en-Halatte, France

ABSTRACT

Methods of systematic analysis of concentrated extracts of French fresh waters are described. First, gas chromatography was used with FID detection and selec-tive detections (S,P compounds by FPD, halogenated compounds by ECD, N compounds by micro- conductimetric detector). The gas Chromatographie results were then used as a guide for GC-MS identification of the pollutants.

By the multidetection approach, more than one hundred compounds of medium vola-tility were identified at concentrations in the range 10~^ - 10 in extracts corresponding approximately to 10" - 10 concentrations in water. Those com-pounds are distributed within 30 different classes. The detected concentrations in the extracts and the frequency of occurrence of these classes are commented : the higher concentrations and frequencies relate to paraffin hydrocarbons, fatty acids, chlorinated aliphatics hydrocarbons, silicones, often more than 10 fo, and then to phtalate esters, chlorinated aromatic and benzene hydrocarbons, often more than 1 fo.

INTRODUCTION

Parallel with work on analytical methodology in connection with the European Project COST 64 b "Analysis of Organic Micropollutants in water", we have been working within a French group dealing mainly with toxicological topics related to "Organic Micropollutants in water". The analytical methods described here as a systematic "Multidetection" approach were used, in support of the work of the French group, to find correlations between composition and toxicological tests : tests being performed on extracts, we focussed on the analysis of extracts, ignoring the methods which apply to water directly.

Tn*e "Multidetection" approach defined in COST 64 b aimed at a restricted number of analytical operations, on water or on water extracts, to obtain a picture as complete as possible of the composition of the bulk of organic pollutants. This approach is opposed to the "selective" approach : one method for a compound or for a class of compounds such as chlorinated pesticides, nitrosamins, organo-mercury compounds From the beginning, the direct coupling of mass spectro-metry with gas chromatography (GC-MS) was thought as the final tool to obtain the multidetection results, bearing in mind that it should only apply to most of the low molecular weight compounds originating from human activities.

87

88 R. Ferrand, M. Mazza and P. Payen

In our systematic multidetection approach we kept Gas Chromatography with selective detection modes (SD-GC) to act as a guide for the GC-MS operation giving final identifications. Mass Spectrometry with direct introduction of the sample into the source (electron impact) was also retained to detect any im-portant polar or medium molecular weight compound which would not have eluted properly from straight-run GC. Liquid chromatography, in HPLC version, was not retained for this systematic work : up to now we use it mainly for preseparation purposes connected with methodology or selective approach but, in systematic screening of extracts, the work after preseparation was thought too heavy.

The chloroform extracts of pollutants were supplied by IRCHA, a laboratory working in the French group : they represent the sum of extracts obtained direc-tly on water at different pH and concentrates eluted from adsorbtion columns.

METHODS OF ANALYSIS

All the gas Chromatographie separations were performed by temperature programming 80 to 280°C, 8°C/mn, on packed columns with silicones 0V1 or 0V101 as station-nary phases.

GC.FID and selective detection modes

We use two Tracor MT 550 chromatographs :

A. One is equipped with two columns and two detectors :

a) A column connected to a flame photometric detector (FPD) for sul-phur or phosphorus compounds, with the appropriate interference filter, provided with a flame ionisation channel (FID). The chromatograms are recorded simulta-neously with the FID trace and the S (or P) trace.

b) A column connected to a Nir_ electron capture detector (ECD).

B. The other Chromatograph is equipped with one column connected to a micro-conductimetric Hall detector with its pyrolysis unit giving NH to provide selective detection of nitrogen compounds.

Typical Chromatographie conditions used were :

a) for FID and S (or P) trace. Glass column : 2 m, 6 mm 0D

Stationnary phase : 6 fo 0V1 on chromosorb WHP Carrier gas : Nitrogen at 60 ml/mn Injection : 0,5 to 5 M.

b) for ECD trace. Glass column : 2 m , 6 mm 0D

Stationnary phase : 2 $ 0V1 on chromosorb WHP Carrier gas : Nitrogen at 60 ml/mn Injection : 1 μΐ of diluted hexane solution (after evaporation

of chloroform solvent unsuitable for this detector)

Multidetection Approach 89

c) far nitrogen trace. Glass column : 2 m, 6 mm OD

Stationnary phase : 3 $ 0V1 on chromosorb WHP Carrier gas ; Helium at 60 ml/mn Injection : 1 to 5 ΜΊ of ether solution (after evaporation of

chloroform solvent, unsuitable for this detector because of HC1 formation by pyrolysis).

Typical chromatograms of an extract are given in Fig. 1 to 4.

Chromatographie resolutions (FID-FPD) correspond roughly to about four compounds separated between two consecutive normal paraffins (Kaiser's separation number). For ECD and N trace we obtain a separation number of about 3.

The Koväts^ retention indexes are obtained by comparison with a separated run on a mixture of :&ctrmal paraffins : by experience we obtain a repeatability of 10 to 20 units.

The semi-quantitative data are obtained first in mass units :

1) for FID with three internal standards (normal paraffins C 1 p - C.^ and Cpn) without any correction for response of individual compounds. These standards are also useful to appreciate and correct retention shifts for the measurement of retention indexes.

2) for S, P and N, in mass of element detected, by use of external standards (dibenzothiophene, triphenyl phosphine and diamino diphenyl methane).

3) for ECD, given in Dieldrin equivalent by the use of an external standard.

We rely then on the concentration of extract in the injected solution and on the volume injected to obtain the concentration in the corresponding extract.

GC-MS

We use a double beam mass spectrometer AEI-MS 30 with an electron impact source, at a resolution of about one thousand-at 3s/decade scan speed, with a data system AEI-DS 30.

The Chromatograph is a Varian Aerograph model 1400.

The home-made molecular separator is Biemann-Watson type ; we previously used a silicone membrane type.

To detect any fault in the molecular separator, a FID detector is connectd in parallel at the outlet of the column, with a system of reducing valves to adjust the flows in FID and MS.


sp.reptreq.s x^iuid^uz

Fig. 2 - Example

of GC-FID + FPD (p)

Extract

152

Fig. 1 - Example

of GC-PID + FPL

(s

) Extract

152

Kovats indexes

H 2

000

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3 Fig. 3 - Example of GC - N trace

Extract N° 152

Fig. 4 - Example of GC -

Extract N° 152 ECD trace

)2 R. Ferrand, M. Mazza and P. Payen

Typical Chromatographie conditions were : Glass column : 1.5 m, 3.2 mm OD Stationary phase : 6 $ 0V1 on chromosorb TOP Temperature programming from 80°C to 290°C at 8°c/mn Carrier gas : helium at 40 ml/mn Flow on FID : about 1/10 of the total flow et the outlet.

A typical chromatogram is given in Fig. 5.

'"SoliUhl" Fig . 5 - Example of GC-MS chronatogra^

Ext rac t N° 152

The operator records mass spectra on significant parts spotted on the chromato-gram during the run. Mass fragmentography is only used for selective detection of a chosen compound or a classes of compounds.

The retention indexes are measured, too, by comparison with a separate Chroma-tographie run on normal paraffins in the same conditions.

Direct introduction Mass Spectrometry

An AEI-MS 9 mass spectrometer is used at a resolution of about 10,000 with direct inlet system in the electron impact source. The temperature of the sample

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Fig. 6 - Example of summary card


is programmed from 40°C to 250°C and several spectra are recorded. In this way we detect compounds with medium molecular weight (up to 1,ΟΟθ) or polar compounds such as medium molecular weight silicones, triglycerids, benzene sulfonic acids.

Detection levels

With selective detectors the detection limits of the Chromatographie systems vary from 10 g at best for ECD up to 10"" g for elemental sulphur, phospho-rus or nitrogen : on about 1 mg injections of extracts the detection limit is better than 10" in concentration units, corresponding to about 10 in water. The confidence level of identifications is only based on retention data but it might be enough to draw attention to the possibility of occurrence of a signi-ficant compound to be checked by GC-MS.

To obtain a significant mass spectrum for identification, the detection limit of the GC-MS system depends on GC resolution and may be in the range 1 to 100 ng : with 100 to 500 ^g of extract injected, the concentration detection level will be then between 10 and 10"·^, in good agreement with experience on peaks detec-ted and quantified by GC and then identified by GC-MS.

Presentation of data

The results of the GC determinations, mainly retention indexes and concentrations are put together with the identification by GC-MS. The difficulty arises some-times from the inaccuracy of retention index measurements : here help comes from frequently occurring compounds such as ethyl ester of satured fatty acids (C1p, CAA, C1fi, C18) and di ethyl hexyl phtalate (DOP) ; after identification, these compounds are located on the profiles of the chromatograms (FID and GC-MS), giving references to match the remaining peaks and to correct shifts in the retention index scale.

The data are presented on "Summary-Cards" ("fiche-resume" in French). A fraction of such a card is shown in Fig. 6. On this card, in the upper left corner, we find the identification of the sample by two code numbers, then the amount of extract in the water ("extrait de l'eau") and the concentration of the injected sample ("extrait sec de l'echantillon").

The card comprises 8 columns :

2 - column 1 : retention indexes X 10 , defining and ordinate axis with

retention times of normal paraffins from C to C corresponding to the examined Chromatographie "spectra".

- the 5 following columns (FID,P,S,N,CE) are bar graph presentations of the chromatograms obtained with the different detectors (CE stands for "capture d'electrons" electron capture) with the concentration evaluated in the units given (10~^ and 10" ). At the bottom end of each of these columns we find the sum of concentrations of defined eluted peaks which could be a significant criterion.

- column 7 relates to MS spectra numbering, still in retention index positions.

- the last column on the right gives the identification from the mass spectral data during the GC-MS run.


In the upper right corner we note additionnal comments, here

- significant occurrence of low molecular weight silicone compounds.

- GC-PPD (P detection) : elution profile of technical tricresyl-phosphate evaluated about 4.10

-3 - CG-ECD profile of PCB estimated at about 2.10 with a majority

of 4 to 6 chlorine compounds.

Latest development in methods

Parallel to this systematic screening of extracts we work on methodology in order to apply recent development to next groups of extracts to be compared. Our work in this field is mainly on the use of home-made glass capillary columns : we use pyrex glass with a home-made capillary drawing apparatus.

Kovats indexes Figure 7 illustrates an example of

the gain in resolution obtained (compare to Fig. l) with an FID detector. The separation number goes from about five to about twenty peaks separated between two consecutive normal paraffins. Here we use a pyrex glass column, KOH treated, of 60 m_ in length and about 0.25 mm in diameter with 0V 101 stationary phase, H carrier gas at 4 ml/mn, and tempera-ture programming from 80°C to 290°C at 8°c/mn. The injection is made by splitless technique.

FPD chromatograms are also obtained now with glass capillary columns with a separation number of about 16 and peak areas are obtained by a micro-processor data system.

For ECD chromatograms, because of dead volumes problems, the separa-tion number obtained is not so good but still with a gain of a factor of about 3 with a column 30 m long and 0.3 nim in diameter. Here we use liquid chromatography preseparation by a short column of silica in order to keep compounds up to the polarity

7 - Glass capillary GC-FID just before phtalate esters : the of extract N° 152 Chromatographie system is safer in

use and we avoid the response of non halogenated compounds such as phtalates.

25004

20004

15004

Now we plan transformations in the laboratory to work with glass capillary, columns on nitrogen selective detector and GC-MS.


COMMENTS ON RESULTS OF SYSTEMATIC SCREENING OF EXTRACTS

Out of more than one hundred extracts examined since we have been working in this field, we took an homogeneous collection of 22 samples to illustrate the potentialities of the set of methods : this collection of samples relates to different points along a French river.

So far more than one hundred individual compounds have been identified at con-centrations level between 10 and 10 in extracts. An extract shows from 30 to 60 GC-FID peaks with concentrations higher than 10"^ representing 10 to 50 fo of the dry extract.

For simplification purposes, in wiew of computer retrieval of the results and to find correlations with toxicological tests, we adopted a classification of aquatic pollutants from the literature on this subject. At the beginning we defined 49 classes of pollutants comprising : hydrocarbons (6 classes), N derivatives (9), P derivatives (3), Cl derivatives (β), 0 derivatives (12), S derivatives (3), organometallies (l), surfactants (2) , optical brighteners (l) and miscellaneous (4 classes). This classification is somewhat arbitrary but helpful to get a broad wiew of the situation.

Out of these 49 classes, up to now we detected 30 (27 in the referred collection of 22 samples) and we use this framework to give here a survey of the results. According to the frequency of occurrence in the collection of 22 extracts (for every class we note n meaning class occurring in n extracts) and the evaluated concentrations in the extracts, we may distinguish several cases.

Frequently occurring classes at high concentrations

Up to concentrations 1 $ and down to frequency F we find 8 classes :

1) Paraffin hydrocarbons (F 22) identified by a paraffinic mass spectral background corresponding to isoparaffins, with sometimes defined GC peaks for normal paraffins. They have not been measured by the set of technique used but are certainly present in concentrations around 10 fo or more.

2) Fatty acids (F , 1 to 20 $). In chloroform solution they appear as ethyl esters, certainly owing to the presence of ethanol as a stabilizer of the solvent. We find 23 compounds, esters of acids from C to C comprising common unsatured C Q compounds. 1 o

3) Light chlorinated aliphatics (F2i, 10"3 to 40%. The class is

represented by 11 derivatives (plus isomers^of hydrocarbons from C to C^ with up to 6 Cl. 2 6

4) Silicones (F_ , 1θ" to 30 $). Light oligomers of polymethyl silo-xanes found by GC-MS plus heavier oligomers found by direct inlet-MS.

5) Phtalate esters (F , 10 to 4 $) . With diethyl ester, coming certainly from esterification of the original acid in water during the extraction, we find classical plasticisers butyl, octyl and decyl esters. We often find another derivative, phtalide, a lactone which is kept in the class.

6) Chlorinated ethers (F , 10 to 7 $). With chlorophenyl phenyl ether, we often find relatively light compounds non completely identified,


but appearing, from mass spectral data, as chlorinated ethers.

7) Chlorinated aromatics (F , 10 to 2 fo). We find chlorinated derivatives of :

- benzene from dichloro to hexachloro

- isopropyl benzene, mainly 3 Cl derivatives

- styrene, 7 and 8 Cl derivatives

- naphtalene, 4 to 8 Cl derivatives

8) Benzene hydrocarbons (F , 10 to 1 fo), with :

- alkyl benzene C to C : lighter homologues are certainly lost in the extracts, because of concentration steps.

- styrene with its dimers and trimers

- diphenyl and polyphenyl up to tetraphenyl.

Lower frequency classes with occasionally high concentrations

Again up to 1 fo concentrations, with frequencies from F to F we find 4 classes :

_ -z — p

9) Miscellaneous N derivatives (F , 10 to 10 ) with :

- isocyanates : phenyl and methyl phenyl derivatives

- phenyl urethanes : methyl, dimethyl and trimethyl derivatives

10) Quinones (F , 10 to 10 ) with anthraquinone and benzanthrone _7 __p

11) Amines (F , 10 to 10 ). Up to now, we detect only alkyl anilins derivatives.

_·? _p

12) Nitrogen heterocyclic compounds (F , 10 to 10 ). With alkyl pyridine derivatives and aminobenzimidazole.

High frequency classes with lower concentrations

We find two classes :

13) Phosphate esters (F , 10 to 10 ). We identify triphenyl and biggest GC peaks of tricresyl phosphate at the higher concentration levels in the collection : as their GC-EPD profile seems significant, they are used to evaluate the technical mixtures.

14) PCB's (Fpp> about 10 ). Here too, GC-MS very often identifies 3 to 7 Cl major isomers and the GC-ECD profile is used for quantification.


Lower frequency, lower concentration classes

Here, we find 12 classes :

15) Miscellaneous chlorinated derivatives (F , about 10 ) with chlo-ropicrin, chloracetaldehyde diethyl acetal, ethyl dichlorobenzoate (probably too from the acid originally in water), mono and dichlorophenyl isocyanates, dichloro indole and chlorinated derivatives of oligomers of silicones (3 to 6 Si).

16) Elemental sufphur (F,„, 10"4 to 1(T5) identified as S_ molecule. 1U o

17) Nitro derivatives (P , 10 to 10 ) with alkyl nitrobenzene derivatives.

18) Chlorinated phenols (F , 10 to 10 ) with higher homologues tetra and penta chlorophenol : the lighter derivatives are either too small or eliminated in concentration steps.

19) Ethers (F , 10 to 10 ) mainly high derivatives not fully identified as their chlorinated derivatives (see class β).

20) Phenols (F^, 10 to 10 ) with higher homologues phenyl, diter-butyl and diterbutyl ethyl phenol.

_·ζ

21) Steroids (F_, about 10 ) with cholesterol and cholestanene. 5

_ _ ■*

22) Polycyclic aromatic hydrocarbons (F , about 10 ). We detect only, in this collection, the major components phenanthrene + anthracene and fluo-ranthene pyrene. In other extracts with higher concentrations, we have already detected chrysene isomers and benzopyrene isomers.

23) Aldehydes (P , 10 to 10 ) with benzaldehyde derivatives.

24) Ketones (F , 10 to 10 ) with acetophenone.

25) ABS surfactants (F ) by direct inlet-MS without quantification.

26) Triglycerids (F ) in the same way as class 25.

Classes rarely detected in other extracts -4 With a lower frequency, down to 10 concentrations, we happened to detect 3

other classes :

27) DDT derivatives, with DDT and DDE, generally in too small concen-trations to be detected without any selective approach.

28) Sulphur heterocyclic compounds : benzo and dibenzo thiophene associated with tar pollutants.

29) Nitrile derivatives : benzo nitrile and alkyl derivatives.

Undetected classes

Among the 48 classes selected, up to now 19 were not detected, either because of


too small concentrations or because they are not attainable without selective

approach for extraction and analysis. They comprise :

- 3 classes of hydrocarbons : olefins, cyclanes and cycloolefins.

- 4 classes of nitrogen derivatives : azo and nitroso compounds, nitrosamines and aminoacids,. The sum of nitrogen eluted in the referred collec-tion fluctuates between 10" and 10 5 with up to 37 detected GC-peaks.

- Chlorinated pesticides, other than aromatics have not been identi-fied. The sum of eluted GC-ECD (as dieldrin equivalent) lies between 10 and more than 1 fo with up to 44 detected GC peaks.

- Miscellaneous sulphur derivatives. The sum of eluted GC-FPD sulphur lies between 10 and about 10 with up to 35 detected peaks.

- 2 classes of phosphorus derivatives : phosphorus-sulphur pesticides and miscellaneous P derivatives. The sum of eluted GC-FPD phosphorus lies bet-ween 10""3 and about 10 with up to 40 detected peaks.

- Organometallics : during methodology work we happened to look at artefacts on phenyl mercury acetate eluted mainly as diphenyl.

- Miscellaneous surfactants, other than ABS.

- 4 classes of oxygen compounds : esters (other ethyl esters, phos-phates and phtalates), alcohols, 0 heterocyclic compounds, 0 miscellaneous.

- Organic pigments.

- Optical brighteners.

ACKNOWLEDGEMENTS

Our work in this field was helped by Financial support of French Environment Authorities, especially SPE (Secretariat aux problemes de l'eau). We are as well indebted to EEC and scientists working on European Project COST 64 b, for helpful discussions on analytical methodology, and to the members of the French group "Micropolluants organiques des eaux", for supply of signi-ficant samples.

Volatile Chlorinated Hydrocarbons in Ground and Lake Waters

W. GIGER, E. MOLNAR-KUBICA andS. WAKEHAM

Swiss Federal Institute for Water Resources and Water Pollution Control (EA WAG), CH-8600 Dübendorf, Switzerland

ABSTRACT

Tetrachloroethylene concentrations in ground waters in the Zurich area have been found to range from less than 0.1 to 82 yg/1, with the highest concentrations in an industrial section of Zurich. A subse-quent survey of ground waters in this industrial area found a series of volatile chlorinated hydrocarbons, including 1,1,1-trichloroethane, carbon tetrachloride, 1,2-dichloroethylene, trichloroethylene, and tetrachloroethylene. In all samples, tetrachloroethylene was the do-minant chlorinated compound, with concentrations between 0.4 and 237 yg/1. The most likely source of this ground water contamination is chronic spillage of small quantities of the solvent.

Vertical profiles of several volatile chlorinated hydrocarbons have been measured in Lake Zurich. Tetrachloroethylene (50-120 ng/1) , 1,4-dichlorobenzene (10-35 ng/1), and carbon tetrachloride (20-35 ng/1) were the most abundant compounds. Profiles of tetrachloroethylene and 1,4-dichlorobenzene showed peaks of maximum concentration at a depth of about 10 m. These maxima are believed to represent a tongue of treated sewage effluent. Thus these two persistent chemicals may be useful as microtracers of wastewater movement in lakes.

Quantitative details of the volatile organic compounds in waters such as those studied for this report are highly desirable since these wa-ters are used as sources for public drinking water supplies.

INTRODUCTION

The significance of biologically resistant organic chemicals for en-vironmental considerations has been documented through many studies of petroleum hydrocarbons, organochlorine pesticides, and polychlori-nated biphenyls. More recently, it has become increasingly evident that volatile chlorinated hydrocarbons are ubiquitous trace contami-

101

102 W. Giger, E. Molnar-Kubica and S. Wakeham

nants in the aquatic and atmospheric environment (1,2,3). Most of these compounds are produced industrially and reach natural waters via accidental or intentional release to the environment. Some halo-genated compounds are formed during water chlorination processes (4); others may have biogenic origins (5). Table 1 summarizes the volatile halogenated hydrocarbons often detected in water samples, including those which will be covered in this report.

TABLE 1 Volatile Halogenated Hydrocarbons

Name

Methyl chloride

Methyl iodide

Methylene chloride

Chloroform

Trihalomethanes

Freon-11

Freon-12

Carbon tetrachloride CC1

Boiling

Vinyl chloride

cis-1,2-Dichloro-ethylene

Trichloroethylene

Tetrachloroethylene

Chlorobenzene

1,4-Dichlorobenzene

F o r m u l a

CH3C1

CH3I

CH 2 C1 2

CHC13

' CHBrCl 2

CHBr Cl

CHBr3

s CHIC1 2

CC13F

CC1 2 F 2

cci4

. H C-CC1 3

Cl H W

H H

C l C l

H H

C l C l / C=c

H C l

C l C l c=c

Cl Cl

< o > c l

C l - ( o ) - C l

p o i n t , C

- 2 4 . 2

4 2 . 4

4 0 . 1

6 1 . 7

90 N

120

1 4 9 . 5

132

2 3 . 8

- 2 9 . 8

7 6 . 5

7 4 . 1

- 1 3 . 9

6 0 . 3

8 7 . 0

1 2 1 . 2

1 3 2 . 0

1 7 4 . 0

Formation

1 / B I , B

I

c , 1

> C

I

I

I

I

I

I

I

I

I

I

Use

Chemical intermediate

Chemical intermedi-ate, pharmaceutical

Solvent coolant,che-mical intermediate

Solvent, pharmaceu-ticals

Cl 1,2,4-Dichlorobenzene Cl & Cl 213.5

Propellants, refrigerants

Solvent, chemical intermediate Solvent

Chemical intermediate

Chemical interme-diate, coolant

Solvent, chemical intermediate

Solvent (dry clea-ning,metal degreasing)

Chemical interme-diate, solvent Constituent of house-hold products, herbi-cide, solvent

Insulator

B: Biosynthesis; I: Industrial Manufacturing; C: Water Chlorination

I

1,1,1-Trichloroethane

Volatile Chlorinated Hydrocarbons 103

Very efficient analytical methods for trace analyses of volatile or-ganic compounds in waters have been developed by K. Grob. Enrichment of the volatile trace constituents is performed by closed-loop gas-eous stripping followed by adsorption on a microcharcoal filter (6, 7). Subsequent analyses of the charcoal eluates by high resolution glass capillary gas chromatography ensures the highest possible se-paration, which has proven to be necessary because of the high com-positional complexity of the mixtures found.

The first applications of this methodology (8), for analyses of va-rious not obviously polluted waters in the Zurich area, revealed the presence of a series of volatile chlorinated compounds (carbon tetra-chloride, chlorobenzene, dichlorobenzene, trichlorobenzene), in ad-dition to many alkylated benzenes, aliphatic hydrocarbons, and seve-ral miscellaneous compounds. An investigation of the Glatt River, using the same methodology, provided information on the levels of volatile organic compounds in a small, highly contaminated river (9, 10). Aliphatic, aromatic, and chlorinated hydrocarbons were detected, as well as aldehydes, ketones, and phthalates. Tetrachloroethylene and 1,4-dichlorobenzene were predominant constituents and exhibited significantly different longitudinal profiles; these differences were attributed to dissimilar sources and differing behaviour in the river.

Analyses of tap water at the EAWAG laboratory have been performed using different enrichment techniques including solvent extraction, gaseous stripping, and XAD-adsorption. The result was always a pre-dominance of tetrachloroethylene by a factor of 10 to 100 compared to other volatile constituents. Quantitative determinations revealed te-trachloroethylene levels of 7.4 to 82.5 yg/1. A subsequent investiga-tion of the sources of this tap water led to the discovery of a con-taminated ground water well with levels of tetrachloroethylene ranging between 500 and 1000 yg/1. Through a study of the aquifer upstream from this particular well, the probable center of the contamination was found. At this site, tetrachloroethylene exceeded its water-solu-bility limit (150 mg/1 at 25°C) and a solvent/water emulsion existed (11). It was never really discovered how this solvent reached the aquifer. There seems to be, however, a close link to a dry-cleaning business since the contamination was centered underneath the back-yard of this facility.

In this paper we report on quantitative investigations of volatile chlorinated hydrocarbons in ground waters in the Zurich area and in Lake Zurich waters. Both types of samples were taken from waters which are used as sources for drinking water supplies. One goal of this study was to provide quantitative measurements of these organic micropollutants to enable an assessment of possible health hazards of such trace contaminants. In addition, better insights could be gained into possible sources and fates of these compounds in the aquatic environment.

EXPERIMENTAL

Sampling

Ground water samples were collected in the greater Zurich area from


44 wells, most of which are used as drinking water sources. Prior to sampling, taps at the particular wells were left open for a minimum of five minutes. One liter glass bottles were thoroughly rinsed with the water being sampled, filled completely, and closed with glass stoppers without leaving a head space volume. Samples were trans-ferred as soon as possible after collection to a cold room (4 C) for storage and were subsequently worked up within about 4 8 hours.

To determine the distribution of volatile chlorinated hydrocarbons in Lake Zurich water, a series of samples were taken during the summer stratification (July, 1977) along a vertical profile at the deepest point in the lake (136 mj off the town of Thalwil). Replicate samples (3 or 4) were collected at each of 15 depths to check on the natural variation in chlorinated hydrocarbon concentrations and sampling errors. A second profile consisting of only four depths was collected at a shallower point (60 m; off Zollikon, close to the raw water in-take of the Zurich Waterworks) one week later for comparison. In both cases, a standard 2 liter metal sampler suspended from a hydrographic wire was used. The water samples were carefully transferred to glass bottles, closed as discussed above, and immediately returned to the laboratory for analysis.

Enrichment

In the preliminary investigation of tetrachloroethylene in ground waters in Zurich, 1 liter samples were extracted three times with 5 ml of pentane by vigorous manual shaking. After adding 1-10 yg of 1-chlorohexane as internal standard, the combined extracts were con-centrated to about 0.2 ml in a Kuderna-Danish evaporator and analyzed by gas chromatography as described below. The quantitative detection limit for tetrachloroethylene by this method was 0.1 yg/1 H2O, with a relative standard deviation of better than 5 %.

Volatile organic constituents in the later study of Zurich industrial area ground water and Lake Zurich water were enriched by the closed-loop gaseous stripping/adsorption/elution procedure developed by Grob (6,7). The high concentration factor (1:10") of this method, coupled with the absence of evaporation steps, permits the analysis of low boiling components at the ng/1 level. For this report, 80 ng each of 1-chlorohexane and 1-chlorooctane internal standards were added to 1 liter water samples, which were then stripped for 90 minutes at 30°C. The volatile components thus liberated from the water were trapped by adsorption onto small activated charcoal filters (1.5 mg of charcoal). The filters were then extracted with a total of about 25 μΐ CS2, and the extracts subjected to gas chromatography. Relative standard deviations for the method based on replicate analyses of spiked (1-chlorohexane and 1-chlorooctane) water samples were deter-mined to be ± 10 % at the 80 ng/1 level.

Gas Chromatography

Qualitative and quantitative determinations of individual volatile components were carried out using a Carlo Erba gas Chromatograph


(Fractovap Model GI) equipped with a flame ionization detector, a Grob-type injector (Model 76, from Brechbühler, Urdorf, Switzerland), and a glass capillary column. Selected samples were also analyzed using dual flame ionization/electron capture detection. Optimum re-solution of the complex mixture of volatiles was achieved using glass capillaries (50,57m x0.3 mm ID) coated with UCON LB 550 according to the barium carbonate procedure described by Grob et al. (12).

Aliquots of 1-2 yl of the CS2 solution were injected onto the column at ambient temperature without stream splitting. After 30 sec, the split valve was opened, allowing the septum and injection port to be purged with carrier gas at several times the flow through the column. Subsequent to the elution of the solvent, the oven temperature was raised from ambient to 150°C at a rate of 2-3 C/min. A flow rate of hy-drogen carrier was maintained at approximately 2 ml/min. Further de-tails of our GC procedure have been outlined elsewhere (13).

Integration of the gas Chromatographie peak areas was performed elec-tronically by a digital integrator (Minigrator, Spectra Physics). Concentrations of several chlorinated hydrocarbons were calculated by comparing peak areas of the compounds of interest with the area of the 1-chlorohexane internal standard and by applying the appropriate predetermined response factors.

Gas Chromatography/Mass Spectrometry

A Finnigan GC/MS system (Model 1015 D) coupled to an on-line computer (Model 6000) was used for mass spectrometric identifications and mass specific detection of selected samples. The UCON LB 550-coated column was directly coupled to the mass spectrometer by means of a platinum capillary. The carrier gas used was helium; an ionizing electron ener-gy of 70 eV was employed.

GROUND WATERS

Motivation

The first part of the study reported here was motivated by the ubi-quitous occurrence of tetrachloroethylene as the major constituent of the volatile organic traces in many natural waters (2,3,5, 8-10) and by the thoroughly investigated case of a locally contaminated aquifer (11) . Furthermore, ground waters are important water reser-voirs for water supplies in Switzerland and in many parts of the world. Therefore, a detailed quantitative knowledge of their organic constituents is desirable. Volatile, biologically refractory consti-tuents are of more concern for ground waters than for surface waters because their elimination through mass transfer to the atmosphere is limited in subsurface waters compared to surface waters. It was in-tended initially to answer two questions: - What are the ubiquitous levels (backgrounds) of tetrachloroethylene in ground waters? - Can elevated levels, if found, be related to a certain type of surface environment (housing, industry, traffic, etc.), or can particular origins be discovered? In addition, the possible occurrence of other


volatile chlorinated hydrocarbons should be investigated both quali-tatively and quantitatively. Conclusions on sources and fate of these compounds could then be drawn.

Results and Discussion

In Fig. 1, a map of the northwestern section of the City of Zurich is presented. The system of surface waters is shown on the map: The Lim-mat River flowing out from Lake Zurich and the Sihl River flowing in-to the Limmat. The flow of the underlying ground water follows first the Sihl and then turns to northwest on the orographically left side of the Limmat. At short distances from the river, an appreciable part of the subsurface water is fed by bank-filtrated river waters. The aquifer delivers water to a large number of water supplies, particu-larly in the Zurich area.

In a first study, ground water samples were collected at locations 1 to 14. These wells are either regularly pumped and used for water supplies or are intended for use in emergency situations. The levels of tetrachloroethylene of these fourteen stations were surveyed by applying the pentane extraction method. The resulting data are pre-sented in Table 2.

TABLE 2 Concentrations of Tetrachloroethylene in Ground Waters of Zurich

No. Name Sampling date, Tetrachloroethylene, 1976 yg/1

1 2 3 4 5 6 7 8 9 10 11 12 13 14

Lochergut Molkerei Migros Schlachthof Kehrichtverbrennung COOP Mühle Löwenbräu Schütze Steinfels Escher Wyss Carba Schöller Hardhof Tüffenwies

July 28 July 28 July 28 July 6 July 28 August 18 August 18 July 28 July 6 July 6 August 18 August 18 March 2-5, July '. July 20

< 0.1 < 0.1 0.6

< 0.1 2.4 0.2 0.4 81.6 4.7 14.8 0.5 0.2

L9 1.7a) 1.5

Numbers refer to Fig. 1

a) Mean of five determinations, observed range: 1.4 - 2.2 yg/1.


Fig. 1. Sampling locations of ground waters in Zurich.

The numbers refer to Tables 2 and 3. The arrows represent the direction of ground water flow.


At three stations (No. 1,2, and 4), the tetrachloroethylene concen-trations were below the detection limit of the technique used (0.1 yg/1). The same levels were found for samples from six additio-nal ground water wells and two springs situated outside of Zurich. Four stations located in a relatively small area contained signifi-cantly higher tetrachloroethylene levels (No. 5,8,9, and 10), ranging from 2.4 to 81.6 yg/1. At one site (No. 13), tetrachloroethylene was determined on five different days, revealing an observed concentra-tion range of 1.4 to 2.2 yg/1 and a mean of 1.7 yg/1.

The results of the tetrachloroethylene survey showed that the ubi-quitous background level of tetrachloroethylene in ground waters is below 0.1 yg/1. The pentane extraction procedure did not provide satisfactory background level data. However, this survey did show that in areas with high industrial activities, elevated levels of tetrachloroethylene in ground waters are encountered. In this regard, the area studied in this report is probably not unique, and such in-creased levels of tetrachloroethylene can be expected in other aqui-fers underlying similar environments.

Station 13 is of particular interest because from this location the Zurich Waterworks pump up to 19 Mio m3 per year for the drinking water supply for the Zurich area. Based on a toxicological assess-ment of traces of tetrachloroethylene by Utzinger and Schlatter (14), the detected concentrations of 1 to 3 yg/1 can be taken to be of no toxicological concern. These authors would consider a daily intake of 50 mg tetrachloroethylene still acceptable. Assuming a daily con-sumption of 3 liters of drinking water, this would set the tolerance limit at approximately 15 mg tetrachloroethylene per liter drinking water. This assessment may be somewhat less relevant because of the recent discovery of the carcinogenic activity of tetrachloroethylene (15). However, no dose-effect relationship for the carcinogenic po-tential of tetrachloroethylene can presently be given, and the extra-polation from results of experiments with rodents to man is question-able. Thus, the question of whether the contamination of drinking waters by trace levels of tetrachloroethylene presents a risk for human populations remains unresolved (14).

From a point of view of taste and odor quality of the water, no detri-mental effects are expected at these concentrations because the thre-shold levels are much higher. The reported threshold odor concentra-tions of volatile chlorinated hydrocarbons range between 5 mg/1 (te-trachloroethylene) and 50 mg/1 (carbon tetrachloride and 1,1,1-tri-chloroethane) (16).

In a subsequent investigation, ground water samples were taken from 17 locations within the area where the highest tetrachloroethylene concentrations had been previously found. The inserted map of Fig. 1 shows the 17 sites (No. 15-31) where ground water was collected. These samples were now treated by the closed-loop gaseous stripping method and quantitatively analyzed for five chlorinated volatile hy-drocarbons. Station 14 (Tiiffenwies) was also included in this more detailed investigation. Qualitatively speaking, three types of sam-ples were found and are depicted in Fig. 2 by respective gas chroma-tograms. In addition to the chlorinated hydrocarbons (tetrachloro-


Station 14 (Tuff en wies)

U4iL

20 aromatic hydrocarbons

40 60 80min

100 150

Fig. 2. Gas chromatograms of volatile organic compounds in ground waters of Zurich.

Enrichment through closed-loop gaseous stripping. Numbers refer to identified compounds listed in Table 3.


TABLE 3 L i s t o f I d e n t i f i e d Compounds

Identifications were made by gas chromatography/mass spectrometry and coinjection experiments. Numbers refer to Fig. 2 and 3.

1

2

3

4

5

6

7

8'

9:

10:

: 1,1,1-trichloroethane

: carbon tetrachloride

: cis-1,2-dichloroethylene

: benzene

trichloroethylene

: toluene

tetrachloroethylene

ethylbenzene

p-xylene

m-xylene

11

12

13

14

15

16

17

18

19:

: chlorobenzene

: o-xylene

: indane

1,4-dichlorobenzene

n-nonanal

2-ethyl-l-hexanol

n-decanal

1, 2 ,4-trichlorobenzene

a ,a-dimethylbenzylalcohol (MS evidence only)

ethylene, trichloroethylene, cis-1,2-dichloroethylene, carbon tetra-chloride, and 1,1,1-trichloroethane), a number of other volatile con-stituents could be detected in all samples. Among them are benzene, toluene, ethylbenzene, xylene, indane, n-nonanal, n-decanal, 2-ethyl-l-hexanol, and a,α-dimethylbenzylalcohol. In Table 3, the identified compounds are listed. The gas chromatogram of station 14 (Fig. 2) is representative for samples from most locations (14-16, 18-22, 26-30). In these mixtures, aromatic hydrocarbons other than benzene, toluene and indane are only minor constituents compared to tetrachloroethy-lene and some other volatiles.

The gas chromatograms of stations 17 (Fig. 2) and 24 show a higher abundance of alkylated benzenes and naphthalenes in a boiling range corresponding approximately to No. 2 fuel oil. At stations 23, 25, and 31, however, higher boiling aromatic hydrocarbons (mostly alkyla-ted benzenes and naphthalenes) are predominant. This type of sample is represented in Fig. 2 by the gas chromatogram of a sample from station 23. From these fingerprint-type analyses one can already con-clude that localized contaminations through petroleum-derived hydro-carbons have occurred near these sites. It seems that such pollutions, which might be caused by either continuously leaking sewer lines or by single spills, are not well dispersed in the aquifer. The signifi-cantly different type of fingerprint found at station 23 compared to station 24 (which is qualitatively similar to the pattern for station 17 in Fig. 2), shows that in this small area (these sites are less than 100 m apart), at least two different sources must be responsible for the contamination of the ground water. It should also be pointed out that in these two samples (sites 23 and 24), the volatile chlori-nated hydrocarbons are less abundant than the alkylated aromatic hy-drocarbons mentioned above.

The results of the quantitative determinations of the volatile chlo-rinated hydrocarbons including the respective detection limits attain-able by the applied techniques are presented in Table 4. In all cases

TABLE 4 Concentrations of Volatile Chlorinated Hydrocarbons

in Ground Waters of Zurich Industrial Quarters

No.

14

15

16

17

18

19

20

21

22

23

24

25

26

27

28

29

30

31

Name

Tüffenwies

Löwenbräu

Kaufmann.Verein

Schütze

Rohr 19

Steinfels

Wohlfahrtshaus

Zivilschutz

Garage

Rohr 18

Rohr 17

Spedition

Forschung

Wassertank

MVA

Kesselhaus

Rohr 16

Lagerhaus

Type

S

S

S S

M S S s

s

M

M S s

s s

s

M s

Tetra-

chloro-

ethylene

2.30

0.44

0.65

44.3

236.4

19.9

26.0

188.4

129.3

1.85

1.21

25.3

39.8

134.8

3.81

37.4

20.5

21.7

Trichloro-

ethylene

0.13

0.59

0.63

0.62

0.14

0.64

1.75

0.16

0.17

0.13

1.92

1.71

1.84

0.57

1.17

0.29

cis-1,2-

Carbon tetra-

Dichloro- chloride

ethylene

1,1,1-Tri-

chloroethane

0.04

0.59

1.10

0.47

0.19

0.06

0.41

3.60

04

30

0.02

0.15

0.33

0.05

0.19

05

12

4.80

0.38

0.10

s3 o

o

er

o 2

Detection limits

0.02

0.02

0.02

0.05

0.02

Values are yg per liter water. Open spaces mean below the detection limits.

Samples were collected on February 15, 1977, except station 14, which was

sampled July 4-11, 1977.

S: Water Supply Well, M: Monitoring Well.

0.

0.


tetrachloroethylene was found to be the predominant chlorinated trace component with concentrations ranging from 0.44 to 236.4 yg/1. Tri-chloroethylene follows as second largest constituent except at sta-tions 29 and 31 where 1,1,1-trichloroethane (4.80 yg/1) and carbon tetrachloride (3.60 yg/1) respectively showed extraordinarily high values. Cis-1,2-dichloroethylene was measurable at four stations from 0.04 to 1.10 yg/1. Carbon tetrachloride was found at four other loca-tions from 0.06 to 3.60 yg/1. At twelve stations, however, 1,1,1-tri-chloroethane could be determined (0.02 to 4.80 yg/1). The analyses at Tüffenwies (station 14) were performed on five consecutive days (July 4-8, 1977) and revealed for tetrachloroethylene an observed range from 2.07 to 2.49 yg/1 and a mean of 2.30 yg/1. One special feature was detected in the single sample from July 6 which contained chloro-benzene, 1,4-dichlorobenzene, and 1,2,4-trichlorobenzene in concentra-tions of 0.113, 0.024, and 0.012 yg/1 respectively. These chlorinated benzenes were below the detection limit (0.001 yg/1 or 1 ng/1) in all other ground water samples.

This more detailed study extends our knowledge on the contamination of an aquifer situated underneath a part of the city of Zurich which is characterized by a number of industries of various types and sizes including machinery manufacture, detergent production, dye-works, and food processing. A certain pattern of volatile organic trace con-stituents is found in all samples, even in those with the lowest con-centration levels. The latter are the two stations closest to the Limmat River (No. 15 and 16 at approximately 100 m distance from the river shoreline) and contain no measurable chlorinated C1/C2 hydro-carbons other than 0.44 and 0.65 yg tetrachloroethylene per liter respectively. At these stations it is thought that the bulk of the subsurface water is bank-filtrated from the Limmat River. Station 14 is hydrologically similar to sites 15 and 16 but about 3 km farther downstream. The significantly higher tetrachloroethylene levels at station 14 (2.3 yg/1) could be attributed to higher contamination of the river waters through the impact of the surrounding urban area. This would seem possible in the light of the longitudinal profile for tetrachloroethylene which had been reported from the Glatt River, a smaller Swiss river flowing through a densely populated area. In the Glatt River, tetrachloroethylene concentrations of up to 80 yg/1 had been detected (9,10).

A second possible explanation would be that the aquifer is contamina-ted at a site upstream. In this context, the level of 1.7 yg/1 tetra-chloroethylene at station 13, as determined in the first part of this study (Table 2), is also of interest. Both levels could be related to the same source because it has been shown that contaminations of sub-surface waters by volatile chlorinated hydrocarbons can be very long-lived and also can be transported over long distances (11). In fact, a heavy tetrachloroethylene contamination of the aquifer caused by an old animal carcass rendering company was discovered in 1974 (17). However, a larger number of monitoring points in this particular area would be needed to clarify this question.

Stieglitz and coworkers (18) have studied waters of the Rhine River and bank-filtrated waters from three waterworks on the Rhine. Since they used very similar techniques (closed-loop stripping and glass


capillary gas chromatography), their data can be readily compared to ours. In the bank-filtrates they also found tetrachloroethylene pre-dominating (0.5 - 2.6 yg/1), whereas in the Rhine River chloroform, carbon tetrachloride, chloro-, and dichlorobenzene were more abundant. In one waterwork, they found much higher levels of chloroform, carbon tetrachloride, trichloroethane, tri-, and tetrachloroethylene in the bank-filtrates. No satisfactory explanation for this phenomenon could be given.

The singular event of the higher contents of chlorinated benzenes at station 14 could be attributed to accidentally increased levels in the river waters feeding the aquifer. Again the investigation of the Glatt River provides valuable information, since 1,4-dichlorobenzene was the second most abundant volatile halogenated constituent in that river. It is, however, difficult to be sure of the significance of this event since it was observed in only one sample. It is worthwhile mentioning that in our analyses no halogenated butadienes were de-tected as in the Rhine River and its bank-filtrates (18). This class of compound seems to be absent as a contaminant in the area which we have studied.

One of the aims of this study had been to precisely elucidate the sources of the investigated ground water contaminants in a manner similar to that in the tetrachloroethylene case reported elsewhere (11). Based on the data provided in Table 4, however, no sources can be pinpointed conclusively. Within an area of approximately 400 x400 m, the levels of all measured volatile chlorinated hydrocarbons fluctu-ate in a range covering up to two orders of magnitude. No clear maxi-ma are evident which would allow one to locate particular sources of contamination. Neither are any concentration gradients seen which could be related to the westerly direction of the ground water flow. In addition to the lack of any pattern of absolute concentration change for the area as a whole, relative concentrations of components also vary greatly from sample to sample. The ratio of tetrachloroethy-lene to trichloroethylene, for example, varies, seemingly at random, from 2 to about 4 00. This fact tends to suggest that different pro-cesses are responsible for the occurrence of this pair of similar compounds.

The most likely reason for this wide variation is that the detected ground water pollution is caused by many different individual conta-minations. Since the investigated compounds are organic solvents which are widely used in different industrial processes (Table 1), it is highly probable that small amounts of them may find their way into the subsurface waters at different locations. Once the volatile chlo-rinated hydrocarbons have reached an aquifer, they tend to persist for long time periods as trace contaminants of the particular ground waters. The persistence and poor dispersal of these chlorinated hy-drocarbons is clearly shown by our results. However, to pinpoint source locations and learn more about transport processes for these compounds will require analysis of a large number of samples over a longer time period.


LAKE ZURICH

Setting and Motivation

The high concentration of tetrachloroethylene and several other chlo-rinated C± and C2 hydrocarbons in ground and river waters in the Zu-rich area led to the question of the presence, concentrations, and processes controlling the distribution of these compounds in the waters of Lake Zurich. Since the lake serves at present as a source for about 75% (approximately 75 Mio m /year) of the drinking water for the City of Zurich and surroundings (population about one million), the presence of volatile chlorinated compounds in the lake water could be of health concern.

Lake Zurich is about 40 km in length, with an average width of 2-3 km. The central basin is U-shaped and has a maximum depth of 136 m (mean depth about 30 m). During summer the lake is strongly stratified, while in winter the waters are mixed. The catchment basin around Lake Zurich is an important recreational and living area. In addition, the lake receives treated wastewater effluents from fifteen sewage treat-ment plants serving about 160'000 inhabitants.

Grob and Grob (8) measured concentrations of trichloroethylene and tetrachloroethylene, as well as several chlorinated benzenes, in sur-face and 30 m deep waters of Lake Zurich. Concentrations of these components (ranging from tenth's to hundredth's of yg/1) were much lower than corresponding levels in ground and drinking waters. The 30 m sample generally contained 2-3 times higher concentrations of these chlorinated compounds than the surface sample. While this con-centration variation was tentatively suggested to result from greater breakdown of the materials at the surface, no further information re-lated to possible sources and concentration-controlling processes was presented.

Results and Discussion

A typical gas chromatogram of the volatiles in Lake Zurich water from a depth of 10 m is shown in Fig. 3 (peaks as identified in Table 3). In addition to the chlorinated volatile hydrocarbons, a complex mix-ture of alkylated benzenes, n-aldehydes, and alcohols were always detected. Vertical concentration profiles for four chlorinated CT/C^ hydrocarbons (1,1,1-trichloroethane, carbon tetrachloride, trichloro-ethylene, and tetrachloroethylene), as well as for three chlorinated benzenes (chlorobenzene, 1,4-dichlorobenzene, and 1,2,4-trichloro-benzene), are plotted in Fig. 4. It should be stressed here that the concentrations are in terms of ng/1 (as opposed to yg/1 used in the ground water discussion in the preceding section). Error bars repre-senting the relative standard deviations obtained from three or four replicate analyses at each depth are given for the tetrachloroethyle-ne and 1,4-dichlorobenzene profiles. For the remaining compounds, standard deviations of about 20% or better were usually obtained; however, for the sake of clarity, the error bars were not plotted.

Volatile Chlorinated Hydrocarbons J15

/w^wvJ uP^ IPm}

LAKE ZURICH 10m

l i jU*^ Ϊ

20 40 60 8 0 min

30 50

7 |S

80

%*Λ*4 tyw (I W Fig. 3. Gas chromatograms of volatile organic compounds in Lake

Zurich water (10 m depth) and in the tertiary effluent of the sewage treatment plant at Thalwil.

Enrichment through closed-loop gaseous stripping. Numbers refer to identified compounds listed in Table 3.

CO

NC

EN

TR

AT

IO

N

(ng

/l)

1,4

- D

ichl

orob

enze

ne

°-™

° 1 7

2,4-

Tric

hlor

o-be

nzen

e x

—x

Chl

orob

enze

ne

OQ

rt

> 1-

4 Pi I

14Q

J

//////

//////

//////

//////

//////

//////

//////

///

//////

//////

//////

//

Fig. 4. Vertical profiles of volatile chlorinated hydrocarbons in Lake Zurich.

TOTA

L TE

MPE

RA

TUR

E (°

C)

OXY

GEN

(m

g/l)

PHO

SPH

ATE

(yg

P/l)

UV

254n

rnA

BS

0RB

AN

CE

0 10

20

0

4 8

12 0

80

16

0 0

I 2

3 4

<3

O o

ET

h-

* O

·-{

H

·

Cö

rt

CD

tu

Vi o

O

DJ

l-i er

o 0

Fig. 5. Vertical profiles of temperature, dissolved oxygen, total phosphate-phosphorus,

and UV-absorbance in Lake Zurich.

UV-absorbance values are for 1 m path length.

118 W. Giger, E. Molnar-Kubica and S. Wakeham»

Note that these errors are due to a combination of factors: sampling error, analytical error, and natural inhomogeneity of the lake at the time of sampling. While the results shown here are only for the de-tailed 15-depth profile, the second profile of 4 points was quantita-tively identical. Also presented for general information (Fig. 5) are a profile of water temperature measured at the time of sampling and several other parameters measured by the laboratory of the Zurich Waterworks which are often used to judge water quality (dissolved oxygen, total phosphate, and UV-absorbance at 2 54 nm for a i m path-length as a measure of dissolved organic material).

Several facts become immediately evident from the profiles of the chlorinated volatiles in Lake Zurich. As in the numerous ground wa-ters discussed above, tetrachloroethylene is the dominant chlorinated volatile compound throughout the water column of the lake. Tetrachlo-roethylene concentrations range from about 50 ng/1 at the surface to about 120 ng/1 at 10 m depth. Below this depth, the measured values of tetrachloroethylene level off at about 60-65 ng/1. Concentrations of 1,1,1-trichloroethane, carbon tetrachloride, and trichloroethylene fluctuate around 8 ng/1, 25 ng/1, and 12 ng/1 respectively, but with-out showing any significant maxima or minima. Chloroform (with a de-tection limit of 1 ng/1) was not found in any sample.

It was somewhat surprising to find such low levels of tetrachloroe-thylene in Lake Zurich waters, especially considering that Grob and Grob (8) measured 120 and 420 ng/1 at the surface and 30 m respecti-vely and in light of the high concentrations present in many ground waters. While tetrachloroethylene concentrations in the ground waters discussed above range from < 0.1 to about 230 yg/1, with a rough mean of about 50 yg/1, the lake water we analyzed contained a mean concen-tration of only about 60 ng/1 - some three orders of magnitude less. The other chlorinated C]_/C2 hydrocarbons are also present in the lake water at concentrations significantly lower than those found in ground waters, although the differences often are not so great as for tetrachloroethylene. It is also worth noting here that the ratio of tetrachloroethylene to trichloroethylene in the lake remains relati-vely constant between 4 and 10, again in contrast to the wide vari-ation of this ratio in the ground waters. One may be tempted to suggest that the levels for these compounds in Lake Zurich may re-present some sort of "background" for these materials in "uncontami-nated" waters, although it is most difficult to assess the chronic sources at such low levels.

The observation that several chlorinated benzenes are present at low concentrations at all depths in the lake (Fig. 4) is interesting since these compounds are apparently present in measurable quantities in only one of the ground waters analyzed. Concentrations of chloro-benzene fluctuate around 2 ng/1 with no trends visible; unfortunately these levels are close to the analytical detection limit of 1 ng/1 for chlorobenzene. The profile measured for 1,4-dichlorobenzene is similar to that of tetrachloroethylene (although at lower concentra-tions) with a maximum concentration of 35 ng/1 at 10 m. On the other hand, levels of 1,2,4-trichlorobenzene gradually (but significantly!) increase from about 4 ng/1 at the surface layer to about 12 ng/1 in the bottom waters.


It is possible, from the results presented above, to draw several conclusions about the processes responsible for the observed distri-butions of volatile chlorinated hydrocarbons in Lake Zurich. Because the vertical profiles of tetrachloroethylene and 1,4-dichlorobenzene are similar (i.e. both showing peaks of maximum concentrations at about 10 m depth), it is likely that the concentrations of these two compounds are being influenced by the same processes. Tetrachloroe-thylene is widely used in the Zurich area as a dry-cleaning and metal degreasing solvent, while 1,4-dichlorobenzene is a common household deodorizer. Both compounds are generally thought to be relatively re-sistant to removal in domestic wastewater treatment plants and in fact tetrachloroethylene and 1,4-dichlorobenzene are major components of the volatiles present in effluents released by such plants. Figure 3 shows a gas chromatogram of the volatiles present in a typical ter-tiary effluent from the sewage treatment plant located at Thalwil -the plant nearest to our lake profile sampling station. Clearly, te-trachloroethylene, 1,4-dichlorobenzene, and 1,2,4-trichlorobenzene are the major components. Fifteen water treatment plants discharge effluents into Lake Zurich, and we believe that the concentration maxima for tetrachloroethylene and 1,4-dichlorobenzene at 10 m re-present a tongue of treated wastewater which, having the proper den-sity, remains in the metalimnion. In this regard, by tracing treated effluents discharged to a stratified lake, Bührer and Ambühl (19) have shown that the effluent tends to flow to the depth of its own density. In the case of Lake Zurich, effluents from the treatment plant at Thalwil, for example, are discharged into the lake at a depth of 6 m and, during June - July, 1977, with a temperature of about 16 C (20). Thus one would expect that this water would sink to about 10 m, where the lake temperature is also about 16 C (Fig. 5) and remain there in a layer. In contrast, toluene and other alkylated benzenes are also present in these treated wastewaters, but at lower concentrations than the chlorinated hydrocarbons (Fig. 3), and do not show a 10 m concentration peak in the lake. The alkylated benzenes are likely to be degraded at the metalimnical oxygen minimum zone where intense microbial activity takes place (see oxygen profile, Fig. 5). It is interesting to note that routine analyses of Lake Zurich waters by the laboratory of the Zurich Waterworks apparently give no indica-tion of any effluent layer at 10 m depth (for example see the parame-ters in Fig. 5). If our interpretation that the tetrachloroethylene and 1,4-dichlorobenzene peaks correspond to a tongue of treated wastewater in the lake is correct, then these compounds may be valu-able as low level tracers of effluent movement in lakes.

It is unlikely that the tetrachloroethylene and 1,4-dichlorobenzene peaks at 10 m depth are the result of release of these compounds from organic material being degraded in the metalimnion. Because of the relatively high water-solubility of tetrachloroethylene and the low concentrations observed to be present in surface waters of Lake Zu-rich, little absorption of this compound into the lipid fraction of organisms living in the surface waters is expected. Thus little will be available for release after the organisms die and sink to the me-talimnion where they are decomposed. In addition, if this absorption/ release-upon-decomposition process were important, then one would expect similar behaviour for the alkylbenzenes (i.e. a peak at 10 m). Clearly this is not the case.


Since about 90% of the tetrachloroethylene and other chlorinated cl/c2 hydrocarbons produced annually are thought to be released to the atmosphere (1) , one is tempted to suggest that atmospheric wash-out by rainfall might provide a major source of these compounds to surface waters. It is possible to test this hypothesis because our lake water profile samples were collected the day following a heavy rainstorm. If rainfall were an important input of chlorinated vola-tiles to Lake Zurich, then one would expect elevated levels of these compounds in surface waters compared to the deep waters. However, the opposite is observed: surface waters are depleted in tetrachloroethy-lene relative to waters in the hypolimnion. The absence of high te-trachloroethylene concentrations at the lakes■s surface indicates that atmospheric washout by rain does not result in a large acute in-put of this compound to Lake Zurich, although rainfall likely is a chronic source for the low level background which is present in the waters. This conclusion is again in contrast to the case of toluene and other alkylbenzenes. The alkylated aromatic hydrocarbons do appear to be brought to the lake by rainout of evaporated gasoline and exhaust. Grob and Grob (8) have shown that concentrations of these compounds in surface lake waters drastically increase immedi-ately following a heavy rainfall, while the levels of chlorinated hydrocarbons do not. Our results on the aromatic hydrocarbons show a concentration maximum in the surface waters of Lake Zurich followed by a rapid and regular decrease in concentration with increasing wa-ter depth.

To account for the slight depletion of trichloroethylene and tetra-chloroethylene at the lake surface relative to waters below about 20 m, several processes must be considered. Microbial degradation is an unlikely cause, since the chlorinated hydrocarbons are generally believed not readily altered by bacteria (hence the persistence of these compounds in biologically treated sewage and the presence of concentration maxima in Lake Zurich at the depth of maximal micro-bial activity). Photochemical decomposition of low-molecular-weight chlorohydrocarbons in the aqueous phase is slow (21), although the chlorinated olefins (trichloroethylene and tetrachloroethylene) would be expected to decompose more rapidly than the chlorinated al-kanes (carbon tetrachloride and 1,1,1-trichloroethane). Evaporation of these volatile compounds from surface waters is, however, likely to play a more important role. Dilling et al.(22,23) reported labo-ratory determined evaporation rates for several of the chlorinated C-./C2 volatiles which we measured in Lake Zurich. Under both quiet and stirred conditions, 1,1,1-trichloroethane, trichloroethylene, and tetrachloroethylene were evaporated from water at the level of 1 ppm to the extent of 90% in about 90 min. Evaporation has been used to account for the concentration decrease for tetrachloroethy-lene along a longitudinal profile of the Glatt River downstream from a suspected tetrachloroethylene source (10,11). In addition, an en-vironmental transport model for tetrachloroethylene concludes that, based on a comparison of atmospheric and oceanic concentrations of tetrachloroethylene, net transport must be from the aqueous phase to the atmosphere (1).


CONCLUSIONS

1. Ground waters in the Zurich area have been found to contain traces of a series of volatile chlorinated hydrocarbons, including 1,1,1-tri-chloroethane, carbon tetrachloride, cis-1,2-dichloroethylene, trichlo-roethylene, and tetrachloroethylene. In one case, chlorobenzene, 1,4-dichlorobenzene, and 1,2,4-trichlorobenzene were also detected.

2. The predominant volatile chlorinated hydrocarbon in all of the ground waters analyzed was tetrachloroethylene. Trichloroethylene was most often present as the second most concentrated component.

3. The highest concentrations of these chlorinated hydrocarbons were found in the industrial quarter of Zurich, suggesting several indu-strially-related sources for this contamination. However, because of the large and seemingly random variation in the concentrations of these compounds within the industrial zone and the lack of concentra-tion gradients which could be related to the ground water flow and to known users of these compounds, it was not possible to elucidate indi-vidual sources for the ground water contamination. Rather, we suspect that a number of different sources are contributing one or several different contaminants.

4. The same series of chlorinated C-L/C^ hydrocarbons, as well as the three chlorinated benzenes are present at all depths in Lake Zurich. Tetrachloroethylene is again the major component, although the concen-trations in the lake are about three orders of magnitude lower than those found in Zurich ground waters.

5. Tetrachloroethylene and 1,4-dichlorobenzene were found to have peaks of maximum concentration in the metalimnion of Lake Zurich. Since these two compounds are abundant in treated domestic wastewaters and are apparently rather resistant to microbial degradation, it is believed that these maxima are due to a tongue of treated wastewater which remains at the thermocline.

6. Both ground and lake waters are used as sources of public drinking water supplies in the Zurich area, and the presence of traces of vo-latile chlorinated hydrocarbons in these waters may present a health hazard. The quantitative data presented in this report should be the basis of an assessment of this potential health concern.

ACKNOWLEDGEMENT

This work was funded in part by the Swiss Department of Commerce (Project COST 64b) and the Swiss National Research Foundation. The authors wish to thank the Institute Bachema, Zurich, for assistance in collecting ground water samples. The support of U. Zimmermann and coworkers of the Zurich Waterworks is acknowledged. We thank K. and G. Grob for their advice throughout the project and for supplying the glass capillary gas chromatography columns used. Comments, advice, and criticisms of F. Zürcher are also appreciated.

W. Giger, E. Molnar-Kubica and S. Wakeham

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National Academy of Sciences (U.S.)· Assessing Potential Ocean Pollutants, Washington, D.C., 1975.

Pearson C.R., and G. McConnell. Chlorinated C^ and C2 Hydro-carbons in the Marine Environment, Proc. R. Soc. Lond. B., 189, 305-332 (1975).

Giger W. Inventory of Organic Gases and Volatiles in the Marine Environment. Mar. Chem. 5, (1977), in press.

Rook J.J. Chlorination Reactions of Fulvic Acids in Natural Waters. Environ. Sei. Techn., 11, 478-482 (1977).

Lovelock J.E. Natural Halocarbons in the Air and the Sea. Nature, 256, 193-194 (1975).

Grob K. Organic Substances in Potable Water and its Precursor, Part I. Methods for Their Determination by Gas-Liquid Chro-matography. J. Chromatogr., 84, 255-273 (1973).

Grob K. and F. Zürcher. Stripping of Trace Organic Substances from Water. Equipment and Procedure. J. Chromatogr., 117, 285-294 (1976).

Grob K. and G. Grob. Organic Substances in Potable Water and its Precursor, Part II. Applications in the Area of Zurich. J. Chromatogr., 90, 303-313 (1974).

Zürcher F. and W. Giger. Flüchtige organische Spurenkomponenten in der Glatt. Vom Wasser, 47, 37-55 (1976).

Giger W., M. Reinhard, Ch. Schaffner, and F. Zürcher. Analyses of Organic Constituents in Water by High-Resolution Gas Chromatography in Combination with Specific Detection and Computer-Assisted Mass Spectrometry. In: "Identification and Analysis of Organic Pollutants in Water", L.H. Keith, Ed., Ann Arbor, Science Publishers, 1976, pp 433-452.

Giger W. and E. Molnar-Kubica. Tetrachloroethylene in Contami-nated Ground and Drinking Waters. Bull. Environ. Contam. Toxicol., 19,(1978), in press.

Grob K., G. Grob, and K. Grob Jr. The Barium Carbonate Proce-dure for the Preparation of Glass Capillary Columns; Further Informations and Developments. Chromatographia, 10, 181-187 (1977).

Grob K. The Glass Capillary Column in Gas Chromatography. A Tool and a Technique. Chromatographia, 8,423-433 (1975).

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2 .

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6 .

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1 3 .

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Utzinger R. and C. Schlatter. Toxicity of Trace Amounts of Tetrachloroethylene in Water. Chemosphere, in press.

Toxic Materials News, 4 (9), 60 (1977).

Dietz F. and Y. Traud. Bestimmung niedermolekularer Chlorkohlen-wasserstoffe in Wässern und Schlämmen mittels Gaschromatogra-phie. Vom Wasser, 41, 137-155 (1973).

Zurich Waterworks, Annual Report, 1974.

Stieglitz L., W. Roth, and W. Kühn. Das Verhalten von Organo-halogenverbindungen bei der Trinkwasseraufbereitung. Vom Wasser, 47, 347-377 (1976).

Bührer H. and H. Ambühl. Die Einleitung von gereinigtem Abwasser in Seen. Schweiz.Z.Hydrol., 37, 347-369 (1975).

Bürgi R., private communications.

Hardie D.W.F., Ed. Kirk-Othmer Encyclopedia of Chemical Tech-nology, 2nd ed., Vol. 5. Interscience Publishers, New York, 1964, pp 183-203.

Dilling W.L., N.B. Tefertiller, and G.J. Kallos. Evaporation Rates and Reactivities of Methylene Chloride, Chloroform, 1,1,1-Trichloroethane, Trichloroethylene, Tetrachloroethy-lene, and other Chlorinated Compounds in Dilute Aqueous Solutions. Environ. Sei. Technol., 9, 833-838 (1975).

Dilling W.L. Interphase Transfer Processes. II. Evaporation Rates of Chloro Methanes, Ethanes, Ethylenes, Propanes, and Propylenes from Dilute Aqueous Solutions. Comparison with Theoretical Predictions. Environ. Sei. Technol., 11, 405-409 (1977).

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23.

Occurrence of Carcinogens in Surface Water and Drinking Water

PROF. DR. J. BORNEFF

Hygiene Institut der Johannes Gutenberg Univ. Mainz, Fed. Rep. Germany

Introduction

In the United States every year about 385.000 persons are dying of cancer, in Germany (FRG) 152.000 cases were registered in 1975. Because of the procentual increase of older people this cancer mortality will probably rise within the next years and decades. International research has improved cancer therapy, the final solution of the problem requires even more effects in prevention of these diseases. Primary condition is the knowledge of the morbidity sources. In this respect it is evident, that exogenic and endogenic causes of cancer are existing and there is no doubt, that the exogenic influences are the more important ones. Their elimination or at least reduction is therefore an urgent interest of preventive medicine.

General aspects of carcinogenesis

Before starting a discussion about different harmful substances in water, a short summary should be given concerning the important points of carcinogenesis. Cancer may occur in mankind as well as in animals. All tissues with cell division may be attacked, but there are great differences in the biological and topographical distribution among the organisms. Epidemiological and geographic-pathological studies have shown, that about 80 % of all maligne tumors are the result of external influences. Cancer is not a fatality as it is often looked upon among people. Especially hard to understand is rhe extreme latency period starting with the influence of the harmful substance up to the realisation of the tumor. For example, the arsenic cancer of the vinedressers is known to take up to 35 years of growth.

Besides the time-factor the correlation between concentration of active substances and carcinogenesis is to be pointed out. High doses of conventional toxins as sublimate or cyanides have a fatal outcome ; little or minor doses might be ingested without any damage; some of those do even have a curative effect, e .g . digitalis. Speaking of carcinogenic substances an often repeated small dose can be more dangerous than one high single shot.

Some confusion was brought about by reports, talking of carcinogenesis as a virus-infection. This has actually been proved only for the Burkitt-tumor in Africa. In our areas the same virus is causing the relatively harmless lymphoblastosis. Thus special cofactors must meet in Africa in order to produce malignant growth. At first sight the lack of epidemic spreading seems to exclude an infectious genesis, but we know about tumor-transmission in animals. On the other hand mankind cannot be infected by a tumor sick person, as it is happening in

125

J26 J . Borneff

case of hepatitis or typhoid fever.

Nevertheless tumor-viruses can be of some importance concerning carcinogenesis. According to the oncogene-theory this is taking place as follows. Carcinogenic viruses have penetrated human cells within a very early developmental phase, here they are existing, reproducing themselves in the same cycle as the cell itself. We can easily imagine that al l human beings are infected. Normally no tumor is growing since the host-organism is capable to control the virus activity. Only if some noxiousubs tan ces have taken influence this control-system is breaking down and the virus may evolve its malignant potency. This theory concerning the virus genesis of cancer is not weakening but underscoring the importance of exogenic cancer noxes.

Finally it should be pointed out, that exogenic noxes may act either indirectly or directly. An example for indirect effect is the colon-cancer. Changes in nutrition can build up a new intestinal flora, resulting in different decomposition products of bile acids, which then may cause cancer. It is the question which substances or compounds do endanger mankind. Apart from the Schneeberger lungcancer, described by Paracelsus, but not recognized as a kind of cancer, the first hint was done in 1775, when Percival Pott reported about the scrota I cancer of chimney sweepers. Anyway, the important basic knowledge is coming from the occupational medicine, that is from direct observations of mankind during 200 years. Animal experiments have been started first in our century.

The polycyclic aromatic hydrocarbons (PAH)are representing a very important group of substances. Kennaway synthetized in 1930 1,2,5,6-dibenzoanthracene as the first pure substance. In 1933 the well known 3,4-benzopyrene was isolated out of coal-tar; its constitution has been proved by synthesis. The simple hydrocarbons like benzene or anthracene are inactive on mice-skin, whereas some of the tetra - and pentacyclic aromates are of high carcinogenic potency. These compounds are effective locally that means after injection of some micrograms into the mice-skin, a tumor is growing at this location. In some cases tumors, like adenomas of the lung, are growing also in distant organs. In the meantime the carcino-genesis of benzopyrene has been traced; within the first step it is transformed to 7,8-epoxide, afterwards water is build into the molecule. Finally 2 diol-epoxides are formed of which one of them can attack the DNA and cause the transformation of the cel l .

Among the group of chlorinated hydrocarbons there are some with clear carcinogenic activity. By vocational observations the liver-tumor-causing effect of vinylchloride has been proved. This is also known for the livertoxic 3 or 4-fold chlorinated methanes. Furthermore poly chlorinated biphenyls are strong carcinogenic substances for the liver of animals. The accumulation of bladder-cancers in people employed in the chemical industry was reason to search for carcinogenic aromatic amines. The first acrylamine recognized was

-naph^thylamine, then benzidine and others. Pure aniline is inactive. Furthermore I like to mention radiomimetic substances like ethylimines and epoxides. The group should be precisely analysed with respect to the ozonization of drinking water.

Of utmost importance are nitrosamines. Long term feeding of animals have shown a high number of liver tumors with a great tendency of formation of metastases. Oesophageal tumors and those of the nervous system do occur, too. As in case of the polycyclics an enzymatic transformation of the molecule into an active compound must take place. N-NO-compoundswould be traced in different foodstuffs moreover it/s formation is possible

Occurrence of Carcinogens 127

biologically, e .g . in the stomach in case primary or secondary amines are meeting nitrites. The stomach is containing catalysators like rhodanides and inhibitors like ascorbic acid on the other hand. Of special importance is the fact, that different drugs lead to nitrosoderivates. In this respect our drinking-water is interesting as a contributor of nitrite respectively nitrate. Out of anorganic-chemical carcinogenic substances arsenic and it's compounds have to-be emphasized. A long term ingestion may lead to benigne and maligne tumors of the skin, mainly, and in some cases of liver and lungs, too. Further the asbestosis is complicated by cancer of the lung. The significance of asbestos in beverages and drinking water with respect to stomach-cancer could not be verified until today. Additionally chromium, nickel and beryllium as substances of anorganic nature are discribed as carcinogenics. Oncogenic substances are produced by nature, too. For example cycasine or mainly mycotoxins and some other products of plants are tumor forming agents.

Another group of compounds are called cocancerogenics; those are not capable to cause a maligne transformation, but can promote a carcinogenic process, which has already been initiated. Concerning surface water some of the phenoles ought to be mentioned. These cocarcinogenics probably play an important part within the human oncogenesis. It is advisable to introduce the term of syncarcinogenesis, which means the cooperation of several carcinogenics aiming at the same organ. However not in every case the combination of 2 noxious substances will lead to a stronger effect, as it is the case with the interaction of benzopyrene and benzoanthracene causing an increase of activity.

On the other hand animal experiments have shown controversal results with other substances. The existence of different carcinogenics in air and food-stuffs, for instance smoked meat, has already been subject for studies; drinking water has not been taken into consideration in this respect enough. The origin of such investigations has been the fact that Wedgewood and Cooper analysed the sewage coming our of gasworks in 1954. They detected benzopyrene and the other aromates yet only qualitatively. Our own study, started in 1959, proved that with regard to the aromates, not only an industrial origin has to be taken into consideration, since benzopyrene was also traced in natural water as well as in domestic sewage, which happened to be a rather important source. In order to clear up the role of these compounds concerning human health, epidemiological studies are absolutely necessary. During the Symposion "Drinking Water Quality and Public Health" 1975 in Medmenham, England, we were informed that Dr. Holland and his coworkers have started corresponding investigations. Since this happens to be prospective study, we can expect any results only later on. Substances out of drinking water may influence cancer-mortality anyway, this fact was proved by a study in New Orleans, which has not yet been disproved.

In this town people being served with haloform-containing Mississippiwater, show a higher bladder cancer mortality than persons receiving unobjectionable groundwater. It is to accept that epidemiological studies often are controversial - keep in mind the discussion about the carcinogenicity of fluoridated drinking-water - but we believe objectionable drinking water makes the frequency of cancer rise most likely. Therefore we ought to analyse water and to el iminate high concentrations. At this time we can give informations about polycyclic aromatic hydro carbons, haloforms; PCB's, nitrosamines and certain metals.

128 J . Borneff

Polycyclic aromatic hydrocarbons

At the beginning of our experiments we assumed, that poiycyclics in water either need the presence of fats , oils, detergents, or that they are taken into the water by means of solid material (e .g. soot). Thus the first part of our program was concerned with the exploration of particles. We checked, for instance, filtermud of a lake waterworks and we discovered several aromates including benzopyrene. We found concentrations up to milligrams per kilogram within the mud. Another analysis showed, that phytoplancton from the Lake Constance contained 1/10 th of the filtermud concentrations. So we concluded, that the major portion of the poiycyclics must be adsorbed to the mineral part of the mud and not to the organic one. Therefore we expanded the studies of Blumer about the occurence of poiycyclics in soil. Besides benzopyrene there were about 30 different polycyclic aromates in every soil sample. We continued these analyses until today and we can say regarding the results of other authors too, that the absence of benzopyrene is an exception. Soil missing any civilizing influence contains about 1 - 10 /ug/Kg, fertilized soil about 100/ug/Kg. Considerably higher values were found in areas/exposed to strong pollution by air aerosols. However, benzopyrene is not only traceable in superficial soil stratums, we also found aromates in material of borings foarm a depth of 170 m and about 100 000 years old, that is in ground, surely not charged by civilization. One of the important sources of aromates in rivers and lakes is the street surface runoff. The main charge we found belongs to the sewage input, where the aromates occur in particles or at least are adsorbed to such. A smaller part is dissolved in the water. Dependent on the degree of air pollution precipitations may contain considerable amounts of poiycyclics. In 16 samples of rainwater Zoeteman and coworkers (WRC Papers and Proceedings Dec. 76) found twice a concentration as in water of the river Rhine. According to other investigations the contamination of snow is much higher than the one of rainwater. Therefore the situation turns out to be as follows. Precipitations will be charged primarily by aromates out of the air-aerosole, they come down on different, always contaminated surfaces like roofs, streets, and agricultural soils. Then they are getting as a surface-runoff into rivers and lakes or they are seeping into the ground. Surface waters will additionally be charged by purification plants flowing offs and furthermore by untreated sewage water. Therefore it is not surprising that even unobjectionable groundwaters have a basic contamination with poiycyclics. Additionally we should mention that tap-water is not always equivalent in its content of carcinogenic substances to the original well water. An increase may happen, if the pipes or reservoirs are isolated with tar or bitumen. Yet , the analysed bitumen, used for pipe constructions, contained only some mg/Kg of carcinogenic material. Further tests showed, that elution will happen for several months. Comparisons pointed out, that generally spoken a low contaminated river- or lake-water is containing 5 times more aromates than ground water, while a medium polluted river water is charged about 1 0 - 2 0 times and a sewage water about 10 000 times higher. A discussion about concentrations of poiycyclics requires the nomination of defined substances, since, "polycyclic aromates" is a collective name for at least 40 partly not yet identified substances. Our proposal is to calculate the sum of the concentrations of fluoranthene, 3,4-benzofluoranthene, 3,4-benzopyrene, 1,1 2-benzoperylene, 11,12-benzofluoranthene and Indeno 0 ,2,3-cd)-pyrene. This was accepted by the German legislation. The following contaminations are average values.


ground and unobjectionable lakewater contains 10 /ug/m^, contaminated river and lakewater 100 ,υα/ηι^ heavy polluted river and lakewater 1000 ,ug/mr sewage water up to 100 000 Ajg/m^ and more.

The surfacewater-contamination apparently tends to decrease, at least in the river Rhine. We presume the enforced construction of sewage water plants and the reduction of a i r

p o l l u t i o n to be the reasons for such a development. We a sce r t a ined the following va lues :

. . . , 3

Rhine near Basel Rhine near Bingen Rhine near Cologne Rhine near Emmerich

1 969/70

270 800 - 3600

1000 - 3500 1000-5600

1976/77+)

210 - 260 4 0 - 9 0

150 - 300 320 - 370

+) Investigations been directed by the Ministerium für Landwirtschaft, Weinbau und Umweltschutz, Rhein land-Pfalz.

The correlation of polycyclics in ground-water to sewage water being a sufficient one, these compounds may serve as sewage indicators. It has to be noticed, however, that street-runoff and tar-coal-contamination is included. According to the actual knowledge errors might happen only, if the rise of polycyclics caused by bituminized or tar-isolated pipes is not taken into consideration. Finding out the reduction of polycyclics is completing the criteria for preparing surface water to drinking water. If you receive 10 ,υα/m^ the procedure is satisfying; otherwise additional suitable steps concer-ning purification should be undertaken. The actual standard for polycyclic aromates has been recommended by the WHO at 200/ug/m . These standards are based on our advice from 1 969. The establishment of these standards was done carefully by intention at this time, in order to avoid unjustified complaints. In the meantime further data hive been collected. Zoeteman reported about 13 random samples of tapwater in Rotterdam, The Hague and Dordrecht during 1973 and 1974. He found a mean value of 45 and a maximum of 300 ,υα/πι^. Our latest analyses in 1976/77 with 294 drinking water samples taken out of 33 big and small waterworks in Rheinland-Pfalz (mostly groundwater supplies) resulted in a mean value of 15 /ug/m . 90 % contained 1 - 25 ,υα/m , 9 % 26 -100 /ug/m^ and 1 % more than 100 ,ug/m3. The maximum value was 247 .ug/m^. None of the samples exceeded the standard of 250 ,υα/m^. Therefore we believe ir to be acceptable to reduce the standard to 100 ,ug/m . '

All people are supposed to be served with drinking water, poor with carcinogenics. Those few cases with higher concentrations need further cleaning methods. The technics are at disposition, there is no need for special development. The highest efficacy is known for the ozonization with following char-coalfiltration. The introduction of such methods is the more important, since carcinogenics of different chemical constitution like PCB's or haloforms can be eliminated also.

Polycyclics in .ug

130 J . Borneff

Volatile chlorinated Hydrocarbons

To this group are belonging chloroforme, carbontetrachloride, 1,2-dichlorethane, vinylchloride, trichlor- ethylene and other compounds, too. In occupational medicine the toxicity of those substances is well known; inebriations, anesthesiae, angiocardiopathy and hepatorenal symptoms can be induced. Tracing these substances in the surface water new problems were shown. The above mentioned observations in New Orleans and analytical studies in the US and the Netherlands were the starting point. In New Orleans the analyses of drinking water resulted in values of chloroforme within the lower ppb-range, i . e . /Ug / I , which means a 1000 times higher concentration in comparison to an artesian wel l , ' exajnined in parallel. Furthermore the corresponding substances were found in the plasma of test-persons. Analyses of the air excluded breathing as one of the sources. hAy report is supposed to tell something about the water contamination in Germany. Not yet published analyses done by Selenka were concerned with surface and drinking water from different areas. The results point out that in general in the Rhine, Main, Weser, Elbe, and Donau concentrations of organic chlorine ranging from 5 to 25 /ug/l can be found. In areas of mainly industrial character e .g . Basel, Rhine-Main-area, Ruhrarea, the amount of chlorine is partly increasing strongly. As a peak the value of 1300 /Ug/I has been traced in the lower Main. Our special interest is aiming at the analyses of drinking water, examined in parallel. Only in few places the concentrations turned out to be smaller than 0,1 /ug/l . Higher concentrations of 10 - 100 /ug/l were found in places, where maximum values had been traced in the riverwater, too. We cannot just compare these results with those from New Orbans, since they have analysed chloroforme and we have regarded the total organic chlorine. The studies of Selenka, however, do show that the main influence is coming from the industrial sewage. At the same time some working concerning the effectiveness of chlorination of the drinking water has been done. In 1976 the concentration of organic chlorine in the Ruhrriver has been measured over 9 months. The mean value was 4 /ug/ l .

After a slow-sand-filtration and a ground-passage the concentration was reduced by half, in order to rise after a chlorination with 0 , 3 - 0 , 6 mg/l up to the original value. This concentration of 4 - 5 /ug/l can be looked upon as a rather low one. In water with plenty of organic compounds we have to expect worse results. Thus the question about the acceptance of chlorine for drinking water treatment is arising. There is no doubt, that we can't leave off disinfectants completely , and that there is only a few number of them available. In my opinion we should not drop chlorine treatment in favour of ozonization, at least not at this point. Drinking water containing high concentrations of haloforms after chlorination can only be badly treated riverwater. In case all available techniques for purification are used, a security chlorination at the end of the treatment is probably acceptable. Undoubtedly chlorine is more suitable for this than ozone. Furthermore I would like to mention, that no research has been done so far , to clear up the question, if ozonization produces carcinogenic epoxides.

PCB A few years after the beginning of their use PCB's have been described as toxic substances in occupational medicine. Since 1966 they are known as ecological harmful substances. In first case they are toxic for the Iiverfunction especially, small amounts can act as enzyme inductors. Combined with ex -HCB tumors of the liver were caused in animal experiments. Similar to DDT, PCB is stored in the fatty tissue and delivered into the blood-stream in case of hunger or hypothermia. Into the environment PCB's are getting by garbage-products of the


industry, especially by those of waste incherators, which are then blowing out the extremly stable PCB's into the atmosphere almost unchanged. The main part is deposited within the surrounding of the emission source, fine particles are getting into areas far away, too. Therefore it is no wonder, that even well protected lakes for drinking water do show a basic load of PCB's. As an example I would like to demonstrate our monthly analyses of the Lake Zürich. The yearly mean value was 5 n g / l in 1972/73, 4 ng/ l in 1974/75 and 6 ng/l in 1976. During the last year we were working with capillar gas-chromatography and defined 8 different t r i - , tetra, penta- and hexachlorine biphenyles. The range was extremely low, the minimum value 5 /Ug/I , the maximum one 6 /Ug/I . Higher concentrations are to be expected in sewage-contaminated rivers. For instance, we found in the upper Rhine 14 ng / l , in the middle part 28 ng/l and in the Rhine river close to the Netherlands border there were 103 ng/l in average. Counting together the mean values out of 13 German rivers, we are receiving a contamination of about 20 ng/l with a maximum of 300 ng / l . The calculation of the Rhine load per year was 1000 kg PCB's. The Institute of Hygiene of Mainz also did investigations concerning the PCB-content of well-water. Thereby concentrations ranging about 10 ng/l could be traced. Apparently the high stability of PCB is causing a ground water load. According to these data mankind is picking up at least 10 /Ug/ of PCB'S by drinking water per year. The amount seems to be a small one at first sight. However one should not forget about the other sources. Additionally the biological chain is causing an accumulation. In USA 1 mg PCB per kg body fat ha s been traced, in Germany however 6 mg, which is 10 times more than Lindane.

Nitrosamines

Nitrosamines are looked upon as very dangerous carcinogenics, since many compounds out of this group happen to be active, causing -in contrary to polycyc I ics-tumors to a high extend if given orally. Their presence in surface water has been assumed, since the precursors are existent. It has been proved, that microorganisms may synthesize nitrosamines out of those substances. We wanted to find out something about the condition of the river Rhine and the Main and worked out the following method of identification: adsorption of nitrosamines to SiO« in a closed circulation, steamdestiNation with simultaneous extraction7gaschromato-graphy with nitrogenspecific detector. The results were proved by GC-MS-investigations . The detection limit for volatile nitrosamines is 0,1 ,ug/l with this method. We analysed in 1976 24 watersamples of the Rhine, taken from the Altrhein at Ginsheim and the Rhine at Mainz. We could not trace any nitrosamines at a l l . Only in a water sample out of the Main we found a minor amount, closely above the detection limit of the method. Thus we cannot exclude the possibility of nitrosamines occuring in water, but the danger through these substances by way of drinking water can almost be neglected.

Metals and metalloides

The accidents happening in Japan the Minamata- disease and the Itai-ltai Illness as well -have drawn the public attention on the toxic metals. The acute and chronical intoxications have already been known for a long time; e.g. in 1473 Paracelsus had written about the mercury-intoxication. Carcinogenic effects, however, have been found out within our century. The problem is, that many metals are essential trace substances in low concentrations, in higher ones they may have a toxic and even carcinogenic effect, e . g . cobalt. Difficulties are arising by the fact that metals and metalloids cannot easily be tested by animal experiments,

132 J . Borneff

since the acute toxicity is preventing a longtime-application of carcinogenic doses. The best example for this is arsenic. The carcinogenicity of this metalloid for mankind is out of question. According to evaluations the needed dose of 50 g appears to be relatively high. Compared to this value the amounts picked up by drinking water are very small. Within surfacewater there are no considerable concentrations of arsenic; a dangerous effect might be caused only by geological contaminations. Such arsenic deposits we can find in Rheinland-Pfalz, where the Max-Quelle of DUrkheim is containing concentrations up to 17 mg/ l . Such a water should not be used by mankind, which is a matter of course, since the carcinogenic dose would be reached within 3 - 4 years. We did find, however, also in rural drinking water supplies concentrations of 0,05 mg arsenic per liter. Such a water makes us pick up about 1,5 g arsenic within 40 years. The assumed carcinogenic dose will not be reached by far, we should not forget the possibility of syncarcinogenesis in this case, either. Beryllium as a carcinogenic does not play any important role in the gaining of drinking water, all analyses, even in heavily contaminated rivers, turned out to be under 0,0001 mg/ l . In contrary you will find selenium in the river Rhine in relatively high concentration compared with ground water. Selenium has been counted to carcinogenic substances for a long time. Today, however, we would rather consider it to be anticarcinogenic substance. Zinc, too, no longer seems to be an ant icarcinogenic. Chromium apparently is causing lung-cancer only by inhalation, affections by drinking water have not been reported yet. On principle we should state, that for all mentioned carcinogenic substances in water, the epidemiological and the experimental investigations as well are not sufficient by far. We ought to work in this field of science even more intensive in order to serve the health of our populations.

Abstract

The mortality tables are ranging malignant tumors right behind coronary diseases in second place. A reduction of the death rates can only be obtained by exploring all morbidity sources. Seen from the standpoint of preventive medicine it is not practicable, however, to leave off prophylactic treatment until the epidemiological or experimental proof for every reason supposed has been demonstrated. We are convinced that just the tentative diagnosis should be a challenge. According to the actual knowledge there are besides the physical cancer sources and the oncogenic viruses numerous chemical compounds, which may have a carcinogenic effect. These are mainly polycyclic aromates, aromatic amines and azo-compounds, nitrosamines, chlorinated hydrocarbons, arsenic, asbestos and some other anorganic-chemical substances as well as several natural products. The cocarcinogenic substances should not be forgotten, either. Within the drinking water we can trace some of them in very low concentrations. However, it is to emphasize that carcinogenic compounds may often have a stronger effect frequently dispensed than in equal doses given in one shot. During the last years epidemiological studies have clearly demonstrated, that there are correlations between water pollution and different types of cancer. Tracing carcinogenic substances in drinking water should therefore be an urgent interest of the water-hygiene.

The occurence of polycyclic aromates is relatively well known in the meantime. In traces these compounds can be found even in unobjectionable ground water. The charges to be tolerated are ranging from 1 - 25 /Ug/m . According to our latest results the recommended standard of 200 (in FRG250) /ug/m should be reduced to 100 Ajg/m within the next time.

Volatile chlorinated hydrocarbons get into the rivers mainly by industrial sewage. This assertion is based on investigations recently done in Germany. Furthermore drinking waters produced in industrial areas carry a load of haloforms above average. The chlorination of


drinking water is in this respect important only, if the elimination of total organic charge is done insufficiently. There is no need to replace all chlorination plants by ozonization at once. An improvement of cleaning procedures in general isof greater importance. PCB'S can be traced in water up to a range of some nanograms per liter only. Because of the enrichment in fatty tissues also such low concentrations should carefully be regarded. By air pollution also drinking water storage basins will be contaminated and a ground water charge is evident. In contrary nitrosamines are not detectable in effective amounts in ground and surface water.

With regard to the metals and metalloides respectively arsenic pollution can be caused by geological formation ; the other substances out of this group do have little or no importance at a l l . With respect to the carcinogenic situation we would like to emphasize the necessity to intensify this branch of water hygiene.

Occurrence and Origin of Non-Biodegradable Contaminants in Surface Waters of The Netherlands

O. VANDEVELDE

Government Institute for Sewage and Waste Water Treatment (RIZA), Lelystad, The Netherlands

ABSTRACT

In 1975 an inquiry was held dealing with discharges of non-biodegradable matter. In the present paper, the results of this inventory are given.

INTRODUCTION

At a national level a program is drafted for the reduction of surface water pollu-tion by non-biodegradable matter. The first requirement is a good knowledge of the nature and volume of the discharges which are taking place. Up to now, a great deal of attention has been given to discharges from industrial sources, because most available data refer to this category. The reason is that this category is very well covered by the Pollution of Surface Waters Act, which came into force in 1970. According to this Act a permit is required for the discharge of all effluent of pollutants or harmful substances into surface water and special conditions with regard to the kind of treatment as well as the quality and quantity of the effluent can be attached to it. As a result, cleaning up operations can be forced by law. It will be clear that such conditions are senseless for the approach to non-industrial sources like phosphates in detergents or lead in petrol. However, the present legislation does not allow attaching conditions at the composition of products on behalf of the abatement of surface water pollution. It should be noted that in the autumn of 1976 RIZA started a project in order to gain a better insight into the occurrence of non-biodegradable matter in sewage, surface runoff, purification plant effluents and sludge. This project is supposed to supply substantial information about discharges from non-industrial sources, in particular dwellings and road traffic. However, the project is expected to be finished in 1978 or 1979. Therefore, the present paper must be confined to discharges from industrial sources.

PROCEDURE

In the past, some information about discharges of non-biodegradable matter was available; in particular this information applied to the chief direct discharges of industrial effluents into the State Waters. However, in view of the program to be drafted, more information was needed. In 1975 an inquiry was held and the proper authorities (the national government, provinces en water boards) were asked to give the amounts of this matter being discharged by the various branches of industry. In processing the figures a distinction was made between discharges from major industries (a number of chemical and metallurgical companies) and discharges from

135

136 0. van de Velde

the remaining industries (e.g. electroplating, textile industry). As a rule, the former industries are concentrated on certain State managed waters such as the Rhine estuary, The North Sea Canal, the Westerscheldt and the Maas. The latter industries are more common and discharge both into State waters and the remaining waters; in addition a considerable part of this category is connected to a municipal sewerage system and to a sewage treatment plant. This distinction had to be made, because in the Netherlands these major industries are unique of their kind; as a result, a more individual approach to the produced water pollution is required. On the other hand the water pollution produced by the remaining industries is liable to a more collective approach in view of the large number of companies and the greater similarity within the relevant industrial branches.

DISCHARGES FROM MAJOR INDUSTRIES.

As mentioned before the first part of the inventory covers a number of major indus-tries which— as a rule—are concentrated on certain State managed waters. The 46 manufacturing establishments in question can be classified as shown in table 1.

TABLE 1 Classification of major industries.

Industries

Petroleum Ferti1izers Basic chemicals Chemical products 1ron and steel Nonferrous metals

No. of establishments

7 k

22 8 1 k

Table 2 indicates the discharges of a number of non-biodegradable substances from these industries. In spite of a certain biodegradabi1ity cyanide and mineral oil are included. In addition, table 2 shows the situation in 1980, considering the remedial measures to be taken.

TABLE 2 Discharges from major industries in tons per annum.

1 ■ _ - ■ ■ -- . . . . . . .. . . .

Cadmium Mercury Arsenic Chromium Copper Lead Nickel Zinc Cyanide Mineral oil

1975

17 3.2 32 290 19 330 14

1,500 1,000 2,900

1980

17 0.2 6 35 12 25 7 I 50 200

1,100

Occurrence and Origin of Non-Biodegradable Contaminants 137

It appears that substantial reductions are expected. However, it should be noted that the situation with regard to cadmium is rather difficult. The greatest problem is represented by the fertilizer factories which produce phosphoric acid by mixing phosphate rock and sulfuric acid. Phosphate rock contains traces of various pollutants including cadmium; both the produced phosphoric acid and the gypsum waste are polluted in this way. The gypsum waste is either dumped or discharged in surface water. Although the concentration of cadmium in the waste is very low, these factories are responsible for the large quantities of cadmium being discharged into State managed waters. These quantities are considerably larger than the quantities discharged by other major industries. By the use of phosphoric acid derived fertilizers comparable quantities are scattered over farming-lands. The possibility of improving the situation either by using the gypsum in the buil-ding industry, by dumping it on suitable sites or by using phosphate rock with smaller traces of cadmium is now being investigated. The situation with regard to mercury is more favourable. In 1975, the manufactu-ring of pesticides containing mercury was responsible for a discharge of 2 tons of mercury each year. Meanwhile this discharge had extremely been reduced by modifying the production. Another category is the chlorine alkali industry, where the mercury electrolysis method is used in the production of chlorine. There are four such chlorine factories in the Netherlands, producing a total of 3^0,000 tons of-chlorine gas annually. These companies worked out plans to make a drastic reduc-tion in the discharge of effluent containing mercury. The plans are now being carried out in stages. Up to a few years ago, the effluent contained 50 grams of mercury for every ton of chlorine produced; in 1973 this loss had been reduced to an average of 10 grams. Meanwhile the emission figure has been reduced to 1 gram per ton chlorine produced and it is expected to be reduced to below 0.5 gram within a few years. Mercury loss in effluent will than have been reduced to less than }% of the original amount. This reduction is achieved by separating the mercury waste from the main effluent and treating it with specially developed ion changers or by treating it so as to allow the mercury to be separated in a very insoluble form. These remedial measures being completed, it is expected that some other discharges of mercury -insignificant up to now- will assume growing importance. With regard to the manufacture of pesticides (in particular chlorinated hydrocarbons) a special working-group is investigating the industrial discharges of these sub-stances and is making guidelines for this category considering the development of technical means.

DISCHARGES FROM THE REMAINING INDUSTRIES

The second part of the inventory covers the remaining industries. As mentioned before, these industries are marked by the large number of companies and the greater similarity within the relevant industrial branches. Compared with the major industries, the information about the remaining industries is even more limited. The reason for this is that in approaching the water pollution by industry the chief direct discharges have come first. In particular information about discharges from industries connected to the municipal sewerage system - e.g. repairing-shops, printing-offices, laboratories - has been inadequate up to now. As a result only a rough estimation of the discharges from the remaining industries can be given. The total annual discharges of the various non-biodegradable substances as shown in table 3 are based on extrapolations of the available figures with regard to a number of industrial branches. In view of the inevitable inaccuracy, the contributions from these branches are given as multiples of 10%. The situation in 1980 cannot be indicated until the national program for the reduction of surface water pollution by non-biodegradable matter will be there.

138 0. van de Velde

TABLE 3 Estimation of the discharges from the remaining industries in 1975«

Industries

2) Chemicals Coatings Leather tanning Metal-ware Tank cleaning Petroleum storage Textile Miscellaneous

Total (in tons per annum)

discharges in %

Cd

100

10

Hg As Cr

10

30 50

10

2 0

Cu

10 10

60

10 10

110

Pb

30 10

50

10

50

Ni

20

80

120

Zn

10

70

10 10

160

CN

100

15

Mineral] oil

60

500

1) The dashes indicate that small discharges of the substances in question are known.

2) Excluding the industries in table 1.

Table 3 shows a striking contribution of the metal-ware category (including electroplating) for a large number of substances. It should be pointed out that these figures partly refer to industries which are connected to a municipal sewerage system and to a sewage treatment plant. Therefore, some non-biodegradable substances including heavy metals are partly retained as a result of adsorption at sludge. As a matter of fact, this is one of the aspects in the RIZA-project mentioned before which are investigated.

TOTAL INDUSTRIAL DISCHARGES.

By counting up the discharges from the major and remaining industries the total industrial discharges of a number of non-biodegradable substances can be roughly estimated. In table k the results are given. In addition the discharge reductions -as given in table 2- resulting from the remedial measures to be taken by the major industries before 1980 are compared with these total discharges (rounded off at 5%).

Occurrence and Origin of Non-Biodegradable Contaminants 139

TABLE k Total industrial discharges.

Cadmium Mercury Arsenic Chromium Copper Lead Nickel Zinc Cyanide Mineral oil

1975 ( in tons per annum )

major industries

17 3.2

32 290 19

330 14

1,500 1,000 2,900

remaining industries

10

240 110 50

120 160 15

500

total

27 3.2

32 530 130 380 130

1,700 1,000 3,400

discharge reductions by major industries before 1980 in % of total industrial discharges

0% I 95% 80% 45%

5% 80%

5% 85% 80% 55% I

Table 4 shows that the contributions from the major industries are considerable with regard to mercury, arsenic, lead, zinc, cyanide and mineral oil and to a less extent with regard to cadmium and chromium. With regard to copper and nickel the contributions from the remaining industries are dominating. Moreover, it appears from table 4 that the remedial measures to be taken by the major industries will highly affect the total discharge of mercury, arsenic, lead, zinc, cyanide and to a less extent the total discharge of chromium and mineral oi1. However, it should be pointed out that these expectations are vul nerable because they just refer to a few companies.

COMPARISON WITH THE INFLUX VIA THE RHINE AND THE MAAS.

In table 5 the estimated industrial discharges are compared with the amounts of contaminants which flow into the Netherlands via the Rhine and the Maas.

TABLE 5 Total industrial discharges compared with the influx via the Rhine and the Maas in 1975 in tons per annum.

Cadmium Mercury Arsenic Chromium Copper Lead Nickel Zinc Mineral oi1

Total industrial d ischarges

27 3.2 32 530 130 380 130

1,700 3,400

Influx via the Rhine and the Maas

160 25 350

2,300 1,400 1,500 650

10,000 24,000

140 0. van de Velde

It should be noted that the influx of cyanide cannot be given because cyanide was not included in the analysis program. Table 5 shows that with regard to the ratio between the total industrial discharges and the influx via the Rhine and the Maas four groups can be distinguished:

- arsenic, copper K. 10% - mercury, mineral oil 10-15% - cadmium, zinc 15-20% - chromium, lead, nickel 20-25%

A FEW CONCLUSIONS.

On the basis of experience gained so far in the inventory, the following conclusions can be drawn:

- a big effort is needed for a more complete picture of the surface water pollution by non-biodegradable matter; not only industrial but also non-industrial sources have to be investigated;

- the remedial measures being completed by industry, these non-industrial sources (dwellings, traffic, agriculture) will assume growing importance;

- although the industrial pollution is reduced, the figures may rise as our knowledge increases, in particular with regard to non-industrial sources.

Contribution of Different Sources to Contamination of Surface Waters with Specific Persistent Organic Pollutants

A. WAGGOTT and A. B. WHEATLAND

Water Research Centre, Stevenage Laboratory, U.K.

ABSTRACT

Specific persistent pollutants in surface waters are considered in connexion with their sources, production, patterns of use, and amenability to treatment. The value of ascertaining and categorizing organic compounds which may reach surface waters is discussed, knowing the range of industrial raw materials employed and data available on toxicity and biodegradability. Factors affecting the persis-tence of organic pollutants within the aqueous environment are considered. Case studies of identification of the sources of specific organic compounds found in water are presented.

INTRODUCTION

Over the past 10 to 15 years there has been increasing concern about the number and nature of organic compounds polluting surface waters, mainly because of poten-tial public health risks resulting from indirect re-use of the water but also because of possible adverse effects of the pollutants on aquatic life. In the U.K. as elsewhere, the situation has been aggravated by the increasing need to use low-land rivers as sources of water for public supply in order to meet the growing demand for water. While it is relatively easy to ensure that supplies derived from such sources are free from pathogenic organisms, it is becoming increasingly difficult to guarantee that the finished waters are free of organic compounds which if ingested over a long period of time might be harmful to health. In many cases Water Authorities have insufficient data upon which to base decisions concerning the safe utilisation of surface waters or the safe release to them of sewage and industrial effluent. In order to provide a basis for decisions, many factors having a bearing on the presence, persistence and toxicity of the substances in question must be considered. These include physical, chemical and biochemical properties of actual and of potential pollutants; also the scale and patterns of production and use, the concentrations likely to be discharged, the toxicological properties of the substances (including persistence and bioaccumulation) and possible removal of the substances during processes of treatment available. All these factors will affect the concentration of an organic pollutant finally exist-ent in surface waters and therefore the contribution of a particular source to contamination of water with that pollutant. The above considerations form a vast and far-reaching subject area; it has been necessary therefore to limit this paper to control of water quality in catchments from the point of view of the origins of organic pollutants.

141

142 A. Waggott and A. B. Wheatland

The possible sources of organic pollution of surface waters are illustrated in Fig. 1, where broken lines indicate relatively minor overall contributions and continuous lines the major contributions.

PUBLIC WATER SUPPLY

- Water distribution system -INDUSTRIAL

WATER SUPPLY

DOMESTIC WASTE WATERS

Ί Γ INDUSTRIAL

WASTE WATER

- Discharge to sewers -

S«wa3e___D^_Si°n. treatment land ! I

SLUDGE

I. SEWAGE 2.AGRICULTURAL EFFLUENT DRAINAGE

'AIR .POLLUTION | i . i

- 7-Precipitation

44— 3. URBAN

RUN-OFF

Ί ,-Landfill-i | disposal

INDUSTRIAL SLUDGE

-Industrial waste-water treatment

4.LEACHATEVROM LANDFILLS

5. INDUSTRIAL EFFLUENT

SURFACE WATERS SEA

Fig. 1. Pathways showing different sources of pollution of surface waters with specific organic pollutants

The diagram shows that at any particular location on a river downstream of an industrial community, there may be pollution from industrial waste waters (either treated or untreated) discharged directly to the river, industrial wastes which have received treatment in admixture with sewage, poorly treated domestic sewage, leachate from landfill sites possibly containing toxic industrial wastes, surface-and ground-water flow from agricultural land and urban areas, and unlawful or acci-dental direct discharge of pollutants. The degree of pollution at the location will depend on the proportion of effluent or 'used1 water in the river. It has been estimated(l) that in certain rivers in the U.K. used, or proposed for use, for water supply, the ratio of river flow to effluent flow varies between 68 and 2:1 for river flows exceeded for 95% of the time. However, in only h out of 15 cases are ratios outside the range 5.8 to 2:1. For 16 out of 17 rivers examined which are not used for potable supply the ratio is between 2.1 and 0.83:1 (1). Some data relevant to some of the rivers used for water supply in the U.K. have also been given by Packham(2) and are reproduced in Table 1 .

FACTORS AFFECTING THE OCCURRENCE OF ORGANIC MICROPOLLUTANTS IN SURFACE WATERS

National figures for the volumes of waste waters discharged into non-tidal rivers, and ranges of concentration of pollutants present (Table 2), while having no direct bearing on levels of organic pollution in specific localities, are pertinent in gauging the extent of the problem. Data available for England and Wales(3)

show that in the early seventies approximately 11% (21.0 x 10 m /d) of the average

TABLE 1 Proportion of Sewage Effluent to Total River Flow at Water Supply Abstraction Points*

River

Abstraction

point

River flow (m/s)

Average

Minimum

Flow

of sewage

effluent

(m3/d)(dwf)

Proportion of

effluent to total

river flow

(%

)

Average Maximum

Note

Blackwater

Chelmer

Dee

Derwent

(Yorkshire)

Eden

Esk

Great Ouse

Hull

Lee

Ouse

(Yorkshire)

Severn

St our

Thames

Wharfe

1.1+

1.7

Langford

Rushes Lock

(above Langford)

Chester

35(b)

Elvington

16.3

Bough Beech

Rushwarp

Foxcote

Clapham

Offord

Brownshill

Denver

Hempholme

New Gauge

Chingford

Acomb

1

2 2 3

1+

3 5

^5

0.63(a)

0.95(a)

8.8(c)

1+.8

1.3

0.53

0.26(d)

«3(d)

.7(d)

.9(d)

.8(d)

,1

Λ

Λ

Tewkesbury

Sharpness Canal

Langham

3.0

Buscot

7.1+

Staines

57.6

Sutton Courtenay 26.0

Swinford

13.0

Teddington

76.

k Addingham

1 h.7

11 ,

15.

1 ,

2,

19,

6,

3.

26,

1 . 1 7

7

1 5

6

32(J

0 3

6

8 3 58

0.11+

0.17

0.1+7

0.1U

0.1U

0.009

0.052

0Λ7

1.26

1.37

0.95

0.099

0.59

1 .26

Ο.56

2.1+3

5.32

0.1U7

0.66

5.3

7 1.61

0.78

10.0

0.005

10.0

10.1

1.3

0.85

6.1

0.21+

20(e)

36(e)

We)

U7(e)

2l+(e)

2.1+

19.3

21+.6

1 .23

h.9

8.8

9.3

6.

1 6

.0

13.1

0.01+

21.7

(a) 1 in 10 year driest average flow

17.9

(b) High peak flood flows inflate average

5.1+

2.9

(c) May be regulated at this level for

10.8

extended periods

1.8

(d) River flow exceeded 90? of the time

9.0

3U.I+

1+

7.0

6.9

21 .

0 31

+.0

11 .

1 3

2.8

2

7.8

21

+.3

20

. k

38

.0

0.3

3

(e) Percentage based on (d)

o o

0

c o

c H

* From Packham(2)

1 .

1 .

2.

8.

2.3

3

.9


annual run-off (190 x 10 x m /d) consisted of effluent discharges, including crude and treated sewage, industr ia l waste waters and mine waters. Industr ia l process waters constituted Ί% (13.7 x 10° nß/d) of the annual run-off and, before discharge, approximately 75% of t h i s had received some variety of treatment. Although approximately only 50% by volume of waste waters discharged by industry received treatment at a municipal sewage works, the tendency i s for industry increasingly to discharge i t s effluent to public sewers for biological treatment in admixture with sewage rather than d i rec t ly into r i v e r s . Since biological sewage-treatment processes are also widely employed for treatment of industr ia l effluents before discharge to surface waters, the effectiveness of these processes in destroying or removing individual organic substances from aqueous solution greatly influences the degree of pollution of surface water from a part icular source or use of the substance.

TABLE 2 Estimated Contributions of Different Sources to Pollution of Surface Waters in England and Wales*

E st imat ed c one ent rat ion ranges of pol]

r 0 expressed as (106 m3) (ng/i)

Approximate daily _ _, n _,_ ^ Source discharge r a n g e s o f fllu^tS

c Qx expressed as TOC

Sewage e f f l u e n t ΐ ΐ 1 5 - 4 0

- domestic 7 10-50

- industr ia l 7 t

Industr ial waste waters 7 30-3000 discharged direct ly to surface waters (excluding cooling water)

Average t o t a l annual run-off f 169

- urban run-off 5 5-83

- agricul tural run-off 16U§ 1-7

- leachate from landf i l l s Negligible 30-1700

* Data from references too numerous to quote. fNo data available since industr ia l effluent discharged to sewerage systems is

always treated in admixture with domestic sewage. fExcluding waste waters. § Calculated by difference.

Production and Patterns of Use of Organic Compounds

Generally there is very l i t t l e information on the quanti t ies of individual organic chemicals that are present in discharges to surface waters, but considerations of the various factors which can affect the presence and concentration of the compound can resul t in fa i r ly accurate predictions of i t s pollution potential being made. The pattern of manufacture and use of the compound is of primary concern. Ident i -fication of the secondary users , ei ther industr ia l or domestic, as well as major industr ial producers will indicate whether d is t r ibut ion in the environment is l ikely to be widespread or res t r i c ted to specific l o c a l i t i e s . Data of the type shown l a t e r in Table 3 need to be broken down further into quanti t ies produced for

Contribution of Different Sources 145

different types of consumption, if prediction« of a qualitative or quantitative nature regarding the final distribution of the contaminant in the environment are to be made. An even more detailed analysis of the conditions under which the compound is manufactured and used will allow estimates of quantities likely to be discharged in final effluent to be made by distinguishing those patterns of manu-facture and use where only minute quantities are released, from· those where significant and uncontrolled amounts are discharged. Having employed usage patterns to identify situations of likely discharge, it is also possible to analyse the data in order to establish the periodicity and extent of the discharge. This could vary from irregular point-source discharges to continuous widespread discharge, with many degrees between these two extremes.

At a particular location at a particular time, any of the 5 sources of pollution shown in Fig. 1 may be dominant. In general, however, industry is the principal source of the specific, persistent organic pollutants which are of greatest concern. Some useful data on the patterns of use of some major industrial chemicals are available(U). The major sources of industrial effluents discharged to non-tidal rivers in the U.K. are listed in Table 3 which illustrates the wide variety of industries involved. The table also lists some of the wide-ranging types of organic compounds employed or produced by the industry and which might be present in· effluent discharges. Information on the actual concentrations of individual components identified in industrial effluent is meagre but where they are available, data have been included.

The actual load of polluting material from the listed industries will vary widely and be largely independent of the total volumes of the discharges involved - e.g. generation of electricity will make a relatively small contribution while for other industries, although the volumes of discharge are relatively small, the potential for pollution may be significant. Ranges of chemical oxygen demand typical of various industrial waste waters are given in Table 3 and may be used as an indication of the contribution a particular class of effluent might make to pollution.

Manuals containing information on industrial and domestic usage of organic chemicals(5~9) and commercial formulations(10) list several thousand organic sub-stances. It has been suggested that the number of chemicals available in the U.K. could be of the order of 10 000 to 20 000 and could be increasing by approximately 1000 per year. Thus, although the complexity of the mixture of organic, compounds present in solution in a treated waste water is suspected, its extent can only be estimated. Theoretically at least traces of all of the organic compounds in solution in an untreated waste water will be present in the final effluent produced from it after biological treatment. This situation arises because concentration is most frequently the limiting factor in bacterial activity. In addition, metabolic products produced by microbial activity and condensation products pro-duced by autoxidation reactions may be present. Thus the final effluent is potentially far more complex than the raw water. On average, the total concentra-tion of organic carbon in solution in a well-oxidized sewage effluent has been found to be approximately 20 mg/l. This is roughly equivalent to kö mg of organic matter/l. To date, concentrations of individual compounds have been measured in the microgram to picogram per litre range. Neglecting those compounds present below a concentration of 1 pg/l and neglecting the fraction containing intractable substances of high molecular weight (the so-called humic acids which constitute approximately 60% of the total organic carbon content of such effluents)

it may be estimated that there could be more than 1.6 x 10 individual compounds in an effluent of this type. Comparison of the given figures with the numbers of organic compounds actually identified and listed in various compilations(11-1U) (usually numbering only approximately 1000) might suggest that there is

TABLE 3 Major Industrial Discharges in the U.K.(U) and some Potential Organic Pollutants

Industry

Overall

rate of

discharge

(I03m3/d)

Source of

effluent

Range of

Some organic constituents

COD*

identified

Concentrations

found £n

discharges

(mg/l)t

Electricity

generation

Mining

f

Paper and board

manufacture

1700

Cooling water

900

Quarrying and

mining (other than

coal)

Chemicals

manufacture

530

U20

Mine water

670

Digestion, bleaching, paper-

making processes

2-Methylnaphthalene

triethyl phosphate

84-10 000 Lignins

Polysaccharides

Optical brighteners

Abietic acid

Terpineol isomers

Trichlorophenol

Vanillin

2-Mercaptobenzothiazole

Pesticide and herbicide

production.

Rubber production

Explosives production

Pharmaceutical preparation.

Heavy chemicals production

ND

ND

0.05-3.3

0.08-0.7b

0.11.5

0.07-4Λ13

0.025-0.0351

Dic

amba

D

ich

loro

pro

p M

CPA

b

is-(

2-c

hlo

roet

ho

xy

)eth

er

But

yl

iso

thio

cyan

ate

Ben

zoth

iazo

le

TNT

2,4

-Din

itro

tolu

ene

To

tal

syn

thet

ic

ster

oid

ho

rmon

es

Py

rid

ine

Tri

chlo

roet

han

e is

omer

s T

rich

loro

eth

yle

ne

1.0

0

.5

1.0

1U

0 0

.1-0

.5

0.02

0.

7 19

0 0.1

s

5-17

5Λ

0.

2

TABLE

3

(contd)

Indu

stry

-

Over all

rate of

discharge

(103 m3/d)

Source

of

effluent

Ran

ge

of

COD*

So

me

orga

nic

cons

titu

ents

id

enti

fied

Co

ne e

ntra

t io

ns

foun

d in

di

scha

rges

(m

g/1)

t Ir

on

and

stee

l 2^

0 m

anuf

actu

re

Age

ing,

co

olin

g,

pick

ling

, ga

s-w

ashi

ng

proc

esse

s In

hibi

tors

S

urfa

ctan

ts

Nap

htha

lene

s Ph

enol

s n o

3

Food

processing

230

600-10000

Fatty

acids

Amino

acids

C ar bo hy dr at e s

Coal

mining

1 50

Textiles

manu-

130

facture

(cotton

and

man-made

fibres)

Coal

-washing.

Coking.

Shale

refining

Kiering,

scouring,

bleach-

ing,

dyeing

processes

1+00

-180

0

3,U-Benzopyrene

Resorcinol

Phenol

Aniline

Adiponitrile

Adipic

acid

Caproic

acid

Decanol

Acrylonitrile

Surfactants

Dyestuffs

0.00005-0.63

150

825-2290

5Λ-16.5

320 3.7

220 2.5

100

Engineering

78

Die moulding, stamping,

casting processes

Bact eric ides

Surfactants

Polyglycols

Phenols

Cyanide

TABLE 3

(contd)

Industry

Overall

rate of

discharge

(103 m3/d)

Source of

effluent

Range of

COD*

Some organic constituents

identified

Concentrations

found in

discharges

(mg/l)t

Petroleum

refining

37

Textiles

21

manufacture (wool)

Gas-scrubbing, rinsing,

processes

Scouring, dyeing, moth-

proofing processes

150-800

2000-5000

Naphthalene

1,3-Xylene

Phenols

Hydrocarbons

Mercaptans

Dichlorobenzene isomers

PCB's

Lanoline

Surfactants

Dyestuffs

0.05

3-1.

0 1

.0

0.00

1 -0

.268

b

0.00

1-0.

021b

03

0Q

0Q

O

03

Soap and deter-

17

gent manufacture

Glue and gelatine

17

manufacture

Washing, kneading processes

2700 §

Alkaline washing.

Vapour condensates

Glycerol

Surfactants

Fatty acids

EDTA

Formaldehyde

Methanol

Urea

Gelatine

03

* Data from a large number of sources too numerous to quote.

fND = Not determined.

fDischarge of mine waters from active and derelict mines (inc. coal mines).

§ Laundry waste water, b: After biological treatment, s: In a crude sewage sample.


considerable further scope for studies in these areas.

A selection of compounds from a list of organic chemicals compiled by Thom(l5) from American data|55l6) is given in Table k. The complete listing is given in Appendix A which shows the major industrial organic chemicals other than gaseous hydrocarbons produced in the U.S.A. However, this may be far more widely applic-able. The listing, as well as Table h9 indicates whether the compounds have been found in sewage effluents, surface waters, ground waters and tap waters and, where they are available, gives concentration ranges. The latter are drawn from a compilation(11 ) made by the Water Research Centre as part of an EEC collaborative exercise (COST Project 6Ub) on organic micropollutants in surface waters. The selected data in Table k give some indication of the persistence of each compound.

TABLE h A Selection* of Major Organic Chemicals Produced in the U.S.A.(15) and Concentrations Measured in Various Waters(lT)

Chemical

Production

^10 kg approx.) Concentration "t^g/l) in

1963(5) 1969(16) J J ^ surface water

ground water

t ap water

Acetone Acrylonitrile

Adipic acid Aliphatic alcohols (C6-<W

Alkylamines (C^-Cj 250

U30

200 805a

300* -- U5-90

1.0

P P P

1 .0

P

A n i o n i c s u r f a c t a n t s C r e s o l s

DDT D i a l k y l p h t h a l a t e s

_o (SfejD-Dichlorobenzene E t h y l e n e g l y c o l E t h y l e n e o x i d e Formaldehyde

N a p h t h a l e n e

O l e i c a c i d P h e n o l

S t y r e n e

T e t r a e t h y l l e a d T r i c h l o r o e t h y l e n e Xy lenes

25Ο 60

80 2U5

750 860

1180

23

10 U10

950

23 170 190

1220

27°

5 5 * 1165 15^0

6 8 0 d

7^5 2110 c

270 1U50

700-1250 0 .03-Uo

0 . 0 1 - 2 5 9 55-2000

1 0 . 0

P

0 . 5 - 2 5 0 . 0 3 - 2 0

P

8 . 6 - 9 . 8 P

10-100000 0 . 0 1 - 1 0 0

0 . 0 0 2 - 5 9 . h 5-350

0 . 2 - 1 6 . 7

0 . 1 - 2 0

P 0 . 0 3 - 3 0

0 . 1 3 - 8 0

0 . 8 2 - 1 3 0 0 0 . 2 - 1 0 0

600-10000

P

P

P

P 0 . 6

15-1^0 1U.6

P 0 . 0 2 9 - 5

0 . 5 - 1 . 0 P

0 . 0 0 1 -10

P 0 . 0 1 -

0 . 0 5

0 . 1 - 5 5 0 U-320

* See Appendix A for f u l l l i s t . •f p denotes presence a t unknown concen t r a t ion . a, 1971 ; b , 1962; c , 1970; d, 1968.

P

P

P


Spaces are left where there is no indication in published literature examined that the compound has been found in the particular class of water. Similar percentages (k29 kk and hQ%) of compounds were found in sewage effluent, surface waters, and tap water, respectively and only 13% in ground water. Of the compounds not detected, 32% were not found in any of the water sources. This can in many cases be explained in terms of physical or chemical properties of the compounds concerned (see the following sub-section). Some of the disproportionately high ranges of concentration quoted arise from incidents of severe pollution and are not typical. Others are caused by particular treatment processes, e.g. ozonolysis and chlorin-ation in the production of drinking water producing high levels of haloforms. Slightly less than half of the identifications are quantified {k9%).

As might have been expected, there is no correlation between production figures and absence or presence (at whatever concentration) in effluents or surface waters. Examples of high production figures and absence or low concentrations (acryloni-trile, ethylene glycol, ethylene oxide, formaldehyde) and low production - high concentrations (chlorobenzenes, cresols, DDT, diethyl phthalates) are apparent. One of the principal causes is the pattern of manufacture, application and distri-bution of the organic compound in question.

Persistence of Organic Pollutants

The basic physical, chemical and biochemical properties of an organic compound have an important bearing on the concentration, persistence and distribution of the compound in the environment. For example, a knowledge of the vapour pressure and activity coefficient of a substance in aqueous solution at ambient temperature will indicate whether it is likely to diffuse rapidly from a body of water into the air. Similarly, polarity and molecular size may indicate whether it is likely to be strongly adsorbed on surfaces and whether its molecules could diffuse through cell walls and enter food chains. A knowledge of chemical properties may indicate whether precipitation of the substance can occur and if it is likely to be destroyed by chemical, photochemical or other oxidations in the natural environment Not least in importance is a knowledge of chemical structure and the rate at which the substance can be degraded aerobically (or anaerobically) by micro-organisms. Although many types of biological transformation occur, bacterial activity is the most important. Biological transformations proceed naturally in aqueous ecosystems Such systems, if not over-polluted, are capable of removing organic pollutants, but complete oxygen depletion or the presence of toxic elements can ultimately lead to either the occurrence of anaerobic conditions or completely sterile environments.

Treatment of waste waters at a sewage works typically involves physical separation of settleable solids followed by biological oxidation of organic matter in the supernatant liquor in percolating filters by the activated-sludge process. After a further stage of sedimentation, tertiary treatment, for example by retention in lagoons or by sand filtration, may then be applied. Sludges are usually disposed of by spreading on land and it is possible that residual organic compounds and their metabolites produced in the soil environment may subsequently be leached out by rainfall and re-enter surface or ground waters although the total concentration of organic matter in the latter is generally very low.

Depending on the substrate involved, the mixed and complex communities of micro-organisms present in biological treatment plant or in natural ecosystems may react in several ways. The substrate may be metabolized quickly and effectively, an acclimatization period may be required before the substrate is effectively removed by the bacterial strains present, a longer period of acclimatization may be required before suitable bacterial strains develop, or perhaps even no significant effect is apparent even after prolonged contact. The processes of biodegradation can lead to practically complete mineralization or organic substrates (i.e. com-


plete conversion to carbon dioxide, water, nitrate, sulphate etc) but will commonly involve both oxidation and assimilation into bacterial cells. In some situations simpler or even more complex metabolites may be produced. The physical state of the organic matter is also important since substances in solution are usually more readily assimilated than those in suspension, and finely-divided sus-pensions are more readily attacked than coarse suspensions. Although a particular organic compound might be metabolized in different ways under different conditions, Thorn and Agg(lT) have found it convenient to consider compounds as belonging to one of three groups - namely, those readily removed, those requiring acclimatiza^ tion of the organisms involved in the breakdown, or those resistant to biological degradation. The demarcation for certain compounds is not always clear, since at moderately low concentrations the biota may readily acclimatize to and success-fully break down a specific organic compound. However, at higher concentrations the compound may be toxic and exert a bacteriostatic or even a bactericidal effect. Temperature is sometimes an important factor also since at low tempera-tures (during winter conditions) removal is relatively poor for some compounds during conventional sewage-treatment processes, while at higher temperatures (during summer conditions) treatment becomes much more efficient.

Apart from biological methods there are available a considerable variety of processes used by industry for the removal of potentially toxic organic substan-ces. The effectiveness of these processes for removal of particular compounds will depend on the physical and chemical properties of the compound, and these proper-ties can be employed to predict the contribution of a particular source, after such water treatment, to contamination of surface waters.

A recent review(l7) lists approximately 250 synthetic organic chemicals about which enquiries have been received most frequently by the Information Service on Toxicity and Biodegradability of the Water Research Centre between 1968 and 197^-. It is thought therefore that these compounds might be among the most important from the standpoint of water pollution in the U.K. Of the listed compounds it is anticipated that less than kO will be effectively removed by conventional biologi-cal sewage treatment. Table 5 compares a selection of these data with those drawn from the compilation(l1) made by the Water Research Centre. The full listings are given in Appendices B to D. The presence and levels of the specific compounds in trade wastes, sewages, sewage effluents, surface waters etc are given.

Of the major industrial organic compounds produced in the U.S.A. and shown in Table h9 63% occur among the U.K. data shown in Appendices B to D, illustrating the similarity of potentially polluting compounds produced in industrial countries. Compounds actually identified in any of the three types of water constitute 59% of those listed and, for surface waters alone, 51, ^> and h3% for biodegradable, slowly degradable, and persistent categories, respectively. The fact that these figures run somewhat against the expected trends when considering biodegradation properties of the three groups of compounds might be explained by the facts that those listed in Table 5 under the headings of "persistent" and "slowly degradable" tend to be more exotic, produced in smaller quantities, and released into receiving waters with a greater degree of control because of their known resistance to biological breakdown. The trend may also be affected by other physical or chemical properties; e.g. lack of data may result from general lack of solubility in water as in the case of carboxymethylcellulose, or the unexpected presence of data may result from the effects of bacterial activity as in the case of acetic and prop-ionic acids in sewage effluent. The fact that as many biodegradable compounds (Table 5) have been identified in sewage effluents and surface waters as in industrial effluents, although at reduced concentrations, illustrates that there is a limit to the removal capabilities of bacteria.


TABLE 5 Concentrations(l1 ) of Biodegradable, Slowly Degradable and Persistent Synthetic Organic Chemicals(l7) in Various Waters

Concentration* (μ-g/l)

unemicajL

Biodegradable

Acetic acid Cresols Glycerol Laurie acid Phenol Propionic acid

Slowly degradable;

Acrylamide Alkylbenzenesulphonat es

(straight chain) Biphenyl Dibutyl phthalate Ethylamine 1-Octanol Oleic acid Trichloroethylene Trichlorophenols Xylenols

Persistent:

Industrial effluent

50-735000

500 5-2290000

P

0.2-1+2

19000 80-1+6000 200

25-115 820-138000

Sewage pfflnpnt

190 0.03-1+0

0.1-0.3 0.03-20

70

280 700-12500

1-80 55-250

0.5-25 8.6-9.8

P P

Surfac e waters

2-20 0.01-100

P 0.03-30 0.1-7.0

0.3-1.2 10-100000

0.0I+-1 5-350

P

0.82-1300 0.05-0.2

P

Alkylbenzenesulphonates (branched chain)

Carboxymethylc ellulose Dieldrin EDTA Pentac hlorophenol Polychlorinated biphenyls

lUl00-5800000 0.2-2.5

0.0003-61+. 3 60-1200 0.2-1+ 0.006-200

0.001-22.5 60-1200 0.1-10000 0.00009-3200

* p denotes presence at unknown concentration.

Where concentration data are available, they appear to bear out the classification of Thorn and Agg(l7). Thus, the expected general trend for easily biodegraded compounds (Appendix B) of decreasing concentrations from industrial effluents to sewage effluents to surface waters is apparent. The trend is not apparent for non-biodegradable compounds (Appendix D) where concentrations in all three waters tend to be within the same order of magnitude. Appendix D, listing those compounds requiring an acclimatization period during sewage-works treatment before removal is effected, contains examples of both trends. The unusually high concentration values occasionally recorded for surface waters can be explained on the basis of local pollution incidence. It is generally more illuminating in following these trends to examine the lowest recorded level where ranges of concentration are quot ed.

P P


fBLACK-LISTST OF ORGANIC COMPOUNDS

Compilations of identified pollutants in various waters covering international 11 , 12,1U), national(l3) and local areas(l8) can be used together with other sources of relevant physical, chemical or biochemical properties to draw up short-lists of substances likely to cause concern. This concern may be because of toxicological properties of the pollutant or in respect of its ability to detract aesthetically from the environment or produce nuisance tastes and odours in finished waters. The main strategy is usually to set priorities for a limited available effort by carrying out detailed surveys for the short-listed compounds.

This above approach has obvious advantages. However, in the U.K. where rivers are relatively short in length, catchment areas small and industries localized, a better philosophy is to apply it to particular catchment areas rather than nationally. The Water Research Centre is at present collaborating with the Thames Water Authority who are carrying out this type of study in the Lee Valley catchment area(19). It is a pilot project to develop the methodology for assessing and controlling persistent low-level pollution of rivers. It may be regarded as a more rigorous approach to control of water quality in a catchment in order to safeguard public water supplies.

Attention was given first to continuous sources of pollution such as discharges from industry, domestic premises and agricultural areas, leachate from landfills and run-off from urban and other areas. Industrial discharges to sewers and direct to rivers were examined and industrial firms involved were visited by senior staff of the Authority. These visits allowed a comprehensive list of industrial organic compounds and their estimated rates of discharge in the catchment area to be compiled. Using available data on relevant physical and chemical properties of the compounds in the list, those questionable organic compounds likely to persist at a concentration and be greater than one part in ten thousand million parts of

10 water (1:10 ) were shortlisted. The Water Research Centre is analysing river water at relevant waterworks intakes as well as finished waters in order to measure the concentrations of the suspect compounds and determine whether actual levels agree with estimated values.

The development of comprehensive lists of organic compounds and the derivation of shortlists of compounds of major concern from them in this way, on a local basis, has many advantages. A picture may be built up of the organic compounds discharged to a particular sewage works, giving the ability to plan ahead by gathering data for the relevant substances on biodegradation, toxicity, removal during various treatment processes, effects of shock loads in the situation of accidental dis-charge, etc. Necessary analytical procedures may also be developed and set up at local laboratories. Similar advantages accrue at waterworks treatment plant where a list of organic compounds likely to be present in raw waters may be compiled.

CASE STUDIES

Within the framework of projects carried out for the Department of the Environment at the WRC on organic micropollutants in effluents, surface waters and finished waters, considerable effort is put into tracing the origins of organic pollutants which have been identified and which are of particular interest. Work in this area can involve both desk studies and experimental work.


Desk Studies

Frequently the tracing of the source of an organic pollutant is simply a matter of reference to a manual of industrial usage(5-10) of organic compounds and making comparisons with known industries discharging to the catchment area under study. This approach is relatively straightforward and not very time-consuming._ It has proved effective in tracing the origins of ethylenediaminetetraacetic acid (EDTA) which was found in relatively high concentrations in the course of a survey of the Lee Valley catchment area(20). Figure 2 is a diagram of the River Lee showing the principal sewage-treatment works discharges and the sampling points. Table 6 gives the concentration data obtained.

R.Lee

i~<D Luton STW

Wheathampstead STW

®l

—<D Harpenden

STW

f— -@

Sewage* Treatment Pilot plant atWRC

Hatfield STW

—06)

Rye Meads STW

R.Stor

(0)-H

# producing sewage effluent from sewage of domestic origin

Girling and George V Reservoirs

Fig. 2. Diagram of the River Lee, showing principal sewage-treatment works (STW)

and sampling points

Contr ibut ion of Different Sources 155

TABLE 6 Concentrat ions of EDTA in Samples from R. Lee and in Eff luents from Sewage-Treatment Works Feeding R. Lee(20)

8 aK£8 Sr*l* HHAWUt (see F ig . 2) J F F i r s t survey f Second survey $

1 R ~ 0 * 6 2 E 1200 6kd 3 R 1120 630 4 E ND* 120 5 R ND* U20 6 E 800 200 7 R 970 290 8 R ND* 1U 9 E ND* 1+20 10 R ND* 280 11 R 33 ND 12 R 310 ND 13 E 630 600 Ik R ND* 110 15 R 260 58

* R = river water; E = sewage-treatment works effluent; ND = not determined. t Mean of duplicate determinations. Average deviation of duplicates from mean, Q%. f First survey, 30 and 31 May 197^; second survey, 1 October 197 - ·

It is apparent from the low initial levels of EDTA (0-1 h /jg/l) found in the river system and the high levels (120-1200 μ§/ΐ) found in all sewage-treatment works effluents discharged to the river system, that the latter are the source of the pollution. The fact that there is a ten-fold difference in concentration between the lowest and highest value recorded for sewage effluents indicates that while there is a diffuse source (suggesting domestic use), there might be point sources of significance (suggesting industrial use). Reference to a manual containing information on usage confirms these impressions(21). EDTA is in fact used for a wide range of industrial, domestic, pharmaceutical, and agricultural applications. These include descaling of boilers, leather tanning, textile dyeing, metal cleaning, separation of metals, and as an ingredient to stabilize soaps, germicides, herbi-cides, vegetable oils, beer, wine, carbonated beverages, milk, dressings and sauces, canned fish and shell-fish and frozen-meat products. EDTA is also used to modify the availability of trace metals to plants and animals and in the treatment of metal poisoning and kidney stones in humans. It is also noteworthy that the compound has been recorded to be resistant to biodegradation(22).

From information elicited from manufacturers it was estimated that approximately 12 50 tonnes per year of EDTA (as the free acid) is marketed in the U.K. Assuming that the EDTA is distributed evenly throughout all sewage flow, and applying the estimate that 95% of EDTA used is discharged in effluents, it may be calculated that the average concentration of EDTA in sewage should be 2 50 Mg/l, which is with-in the range actually found.


Experimental Studies

Occasionally organic compounds identified in water samples are not found listed in the manuals of industrial usage or recipes of commercial formulations. In such cases, further detective work and experimental work is required in order to trace the origins of the compound. An example of this occurred during an examination of some of the coloured materials in sewage effluents(23). On passing a sample of domestic sewage effluent through a column of fAmberlitef XAD-2 resin, the head of the column became blue and a resulting methanol extract was green. The spectrum in the visible region of the extract showed a peak maximum at 63*+ nm. The origin of the dye was traced to the sanitary blocks used, in lavatory flush systems, to release to the flush water non-ionic and cationic surfactants (the latter acting as germicides), perfume, and a blue dye giving the water a fresh appearance and indicating by its fading when the block needs to be replaced. The identity of the dye was stated by the manufacturer of such blocks to be CI Acid Blue 1 which is contained in the commercial dye, Lissamine Turquoise VN 150. The identity of the dye was confirmed by comparing the visible spectra of extracts of an XAD-2 column after passage through it of an aqueous solution of Lissamine Turquoise VN 150, a sample of flush water from a lavatory containing a cistern block, and a sample of domestic sewage effluent. Various colorimetric confirmatory tests were also carried out.

The dye, which has the triphenylmethane structure shown in Fig. 3, is non-volatile and polar. Its polarity might be expected to prevent adsorption on the resin, which usually retains the least polar compounds, but in fact the efficiency of recovery using the resin is 80 to 90? Table 7 shows concentrations of the dye measured at various sampling points in the Lee Valley (see Fig. 2). Concentrations found in the River Lee system suggest that the dye is persistent. Other uses which might provide both local and diffuse sources include the dyeing of textiles (both natural and synthetic), printing, dyeing of leather, paper staining, and the coloration of a wide range of material, e.g. soaps, inks, and methylated spirit.

(ψ)ο^( V-c·

N(C,HJ 2"5'2

«? ® N(C 2H 5) 2

Fig. 3. Structure of CI Acid Blue 1 (Calcium di-U-£ U-diethylaminocyclohexa-2,59-dienylidene-(U-diethylaminophenyl)methyl]benzene-1,3-sulphonate)

A corollary to the above situation where the use of a suspect organic compound is known, is the development of an analytical technique and the performance of survey work in order to determine the concentrations entering surface waters. An example of this is work being carried out at the Water Research Centre on the determination of the contraceptive steroid, ethynyl oestradiol, in sewage effluent. Preliminary studies using information from the Family Planning Association on known dosage and excretion rates allowed calculations on the probable degree of dilution in sewage to be made. Using further information on the known biodegradability of contraceptive steroids(2U) it was calculated that ethynyl oestradiol, which is resistant, would be the only compound likely to be discharged to surface waters at a measurable concentration (of the order of 5 ng/l). Preliminary results employ-ing multiple-ion detection gas/liquid chromatography-mass spectrometry indicate


that the concentration of ethynyl oestradiol in a domestic sewage effluent is in fact of this order, showing that, with accurate data, useful predictions on levels of organic compounds being discharged to surface waters can be made.

TABLE 7 CI Acid Blue 1 Concentrations in Samples from R. Lee and in

Effluents from Sewage-Treatment Works Feeding R. Lee(23)

Sampling ς Ί

point . m a Concentration ^g/l)i* (see Fig. 2) ^

3 R 6.0 6 E 8.7 13 E 8.8 15 R 2.5 16 E 12.0

* See footnote to Table 6. t Samples collected 13 Aug. 1975

Investigation of Complaints

The origin of organic pollutants reaching surface waters, and sometimes tap waters, and causing nuisance problems therein is sometimes more easily traced by analysis of suspect industrial effluent discharges - and frequently this is required for proof of origin. The undertaking of this type of investigation is becoming increasingly common at the Water Research Centre.

An example(25) occurred when the Laboratory was recently asked to produce proof of origin of some industrial organic chemicals causing nuisance odour problems at a sewage works and along reaches of the river to which sewage-works effluents were discharged. Taste and odour problems associated with the compounds were also being experienced in waters abstracted in lower reaches of the river for potable supplies. Using head-space analysis and capillary column gas/liquid chromatography-mass spectrometry, the problem-causing compounds were identified in samples of indus-trial effluent, sewage effluent and river water as terpenes (chiefly limonene and citronellol) occurring at concentrations greater than the threshold odour concen-tration. Other, as yet unidentified compounds, were also traced through the system to the finished water. None of the organic compounds in question was present in the river above the point of discharge.

CONCLUSIONS

A comprehensive knowledge of the potential sources of pollution of organic com-pounds is of great importance in establishing origins of pollution in specific cases. Maintaining this knowledge requires a continuous up-dating of information on production of new compounds, new applications, rates of consumption, and patterns of use as well as basic information on relevant physical and chemical properties.

Hard data on concentrations of organic compounds in various typical environmental and process waters are scarce and the analytical techniques generally required to obtain values are relatively complex, difficult and expensive to apply. However,


by applying accurate data on consumption, biodegradation, etc, in a systematic way, valuable estimates of concentrations of organic pollutants reaching surface waters can be provided. Adopting procedures of this type and using existing knowledge of toxicological effects in order to compile 'black-lists1 of compounds is therefore a valuable exercise, particularly for specific river basins, both from the point of view of control of water quality in catchments and because in many cases it allows limited and expensive analytical effort to be focussed on compounds or groups of compounds potentially of the greatest concern.

ACKNOWLEDGEMENT

This paper is reproduced by permission of the Director of the Water Research Centre.

REFERENCES

(1) W. F. Lester. Standards based on the quality of the receiving water. Wat. Pollut. Control 68, 32U (1969).

(2) R. F. Packham. Potable water from sewage effluent. Paper presented at a sym-posium, Sewage Effluent as a Water Resource, 1^-15 Nov., 1973. Inst. Publ. Hlth. Engrs, London (1973).

(3) Ministry of Housing and Local Government and Welsh Office. Taken for granted. Report of the working party on sewage disposal. HMSO, London (1970).

(k) Department of the Environment and Welsh Office. Report of a river pollution survey of England and Wales 1970, Vol. 2. Discharges and forecasts of improvement. HMSO, London (1972).

(5) W. L. Faith, D. B. Keyes and R. L. Clark (1966). Industrial Chemicals, 3rd edn. Wiley, New York.

(6) H. F. Mark, J . J . McKetta and D. F. Othmer (1965), Kirk-Othmer Encyclopedia of Chemical Technology, Interscience, New York.

(7) J. B. Berkowitz, G. R. Schimke and V. R. Valeri. Water pollution potential of manufactured products, catalog section III - chemical ingredients listing. Envir. Prot. Technol. Ser. Rep. EPA-R2-73-179d, April (1973).

(8) F. A. Patty (ed.) (1967) Industrial Hygiene and Toxicology, Vol II. Toxicology. Interscience.

(9) International Technical Information Institute, Japan (1976). Toxic and Hazardous Industrial Chemicals - Safety Manual for Handling and Disposal.

(10) J. B. Berkowitz, G. R. Schimke and V. R. Valeri. Water pollution potential of manufactured products, catalog section II - product listing. Envir. Prot. Technol. Ser. Rep. EPA-R2-73-1 79c, April (1973).

(11) A. Waggott. A comprehensive list of organic pollutants which have been identified in various surface waters, effluent discharges, aquatic animals and plants, and bottom sediments. Report prepared for European Coopera-tion and Coordination in the Field of Scientific and Technical Research -COST Project 6Ub, Analysis of Organic Micropollutants in Water.


Environmental Protection Agency, Water Quality Criteria Data Book, Vol. I. Organic chemical pollution of fresh water. Wat. Pollut. Control Res. Ser. Rep. 18010DPV12/70, Dec. (1970).

G. A. Junk and S. E. Stanley. Organics in drinking water, Part I. Listing of identified chemicals. U.S. Energy Res. Develop. Admin. July (1975).

W. M. Shackelford and L. H. Keith. Frequency of organic compounds identified in water. Envir. Monit. Ser. Rep. ΕΡΑ-6ΟΟΛ-76-Ο62, Dec. (1976).

N. S. Thorn. The significance of synthetic chemicals in rivers used as a source of drinking water. Technical Note No. Ik. Central Water Planning Unit, Reading, U.K. (1976).

J. A. Kent (197*0. Riegel1 s Handbook of Industrial Chemistry, 7th edn. Van Nostrand Reinhold, New York.

N. S. Thorn and A. R. Agg. The breakdown of synthetic organic compounds in biological processes. Proc. R. Soc. Lond. B. 189, 3^7 (1975).

S. Torrance, Thames Water Authority, Scientific Branch. Private communi-cation.

H. Fish and S. Torrance. Water. J. Nat. Wat. Council No. 15, 15 (1977).

J. Gardiner. Complexation of trace metals by ethylenediaminetetraacetic acid (EDTA) in natural waters. Wat. Res. 10, 507 (1976).

H. F. Mark, J. J. McKetta and D. F. Othmer (1965). Op. cit. Ref. 6, Vol. 6, pp. 21-21+.

R. L. Bunch and M. B. Ettinger. Biodegradability of potential organic sub-stitutes for phosphates. Proc. 22nd ind. Waste Conf., Purdue Univ., Engng Extn Ser. No. 129, 393 (1967).

J. Gardiner. Water Research Centre Technical Report. In preparation.

H. H. Tabak and R. L. Bunch, Steroid hormones as water pollutants. Proc. Symp. Developments in industrial microbiology, Washington, 367-376 (1970).

A. Waggott, J. Gardiner, W. J. Reid and I. W. Davies, Water Research Centre Enquiry Reports Nos. U72 of Feb. 1977 and 52U of July 1977-

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APPENDIX A: MAJOR ORGANIC CHEMICALS PRODUCED IN THE U.S.A.(15) AND

CONCENTRATIONS MEASURED IN VARIOUS WATERS(11)

Production (10" kg approx.

Concentration (μβ/l) in

Chemical

Ac etaldehyde Acetanilide Acetic acid* Acetic anhydride Acetone Acrylonitrile

Adipic acid Aliphatic alcohols

(C6-C18> Alkylamines (C -C )

Alky lb en z en e sul phona t e s Alkyl sulphates 1-Amino-2-propanol

(isopropanolamine)

1963(5) 1969(16) s e w a S e surface ground tap effluent water water water

0.1

P P 1 .0 P

360 1.8 500 590 U30 200

"b 300 -

25Ο

250

750 1.8

1600 760

-805 a

-1+5-90

-

t

P

190

P

P

700-12 :

2-120

1 .0 P

P P

P P t 700-12500 10-10000 600-10000 15-1 Uo t p p

A n i l i n e A n t h r a q u i n o n e A s p i r i n

( a c e t y l s a l i c y l i c a c i d ) Benza ldehyde

Benzene

Benzo ic a c i d Benzy l c h l o r i d e 1 - & 2 - B u t a n o l 2 -Bu tanone B u t y l a c e t a t e t e r t - B u t y l p h e n o l s e -Capro lac ta rn Carbon t e t r a c h l o r i d e

Cärbony l c h l o r i d e ( p h o s g e n e )

C a r b o x y m e t h y l c e l l u l o s e s C e l l u l o s e a c e t a t e C h l o r a l C h l o r o a c e t i c a c i d C h l o r o b e n z e n e C h i o r 0 e t h a n e Chloroform Ch lo rome thane ( m e t h y l

c h l o r i d e )

C i t r i c a c i d C r e s o l s

70 1.8

13

1.8

2160

7 -

270 110

50 --

235

75.

20 270

36 25

235 270

50 50

50 60

180* --

-U080a

---

220 -16 -

Uoo

-

-----

310 100 180

6oc

-

P

P

P

P P

P

P P P P

0.3-Uo

9.8 2 . 0

0 . 1 6 - 2 0 0

P

5 .0 P

0 . 3 - 1 7 0 0

0 . 1 6 - 1 7 0

P P

0 . 0 1 - 1 0 0

0.03

0 . 0 0 0 1 -160 B

P

P P

0 . 5

P 0 . 0 0 5 -

2600

0 . 0 1 - 5

0 . 1 - 1 .0 P P P

U.G

P

P

P

P


Chemical

Production (10° kg approx.)

Concentration (μβ/ΐ) in

ιο^ίθ io£oH^ sewage surface ground tap 1963(5) 1969(16) e f f l u e n t w a t e r w t e r v a t e r

Crotonaldehyde Cyclohexane 2,U-D 21 DDT 80 1-Decanol 27 Detergents : anionic -

nonionic cationic

Dialkyl phthalates 2U5 1 ,2-Dibromoethane 2,6-Di-tert-butyl-p-cresol

£~ &£-Dichlorobenzene Dichlorodifluoromethane Other chlorofluoroalkanes (fluorocarbons)

1,2-Dichloroethane Dichloromethane Diethylene glycol Diethyl ether 1+0 Dimethylac etamide D imet hyl f ormam i d e Ethanol 235 Ethanolamines (mono-, di- 70

tri-) Ethyl acetate 52 Ethylbenzene

Ethylenediamine Ethylene glycol 750 Ethylene oxide 860 Fatty acids (incl. oleic, palmitic and stearic)

Formaldehyde 1180 Formic acid 8

Glycerine 135 1 ^^^-Hexachloro- 5

cyclo hexane (BHC) Hexamethylenediamine -

Hexamethylenetetramine 18

1100

7d

P <1-360

P P

p .01-0.91 27 a 0.01-259 0.002-59.k p

P 1220 700-12500 10-100000 6θΟ-10000 15-1U0

U0 530 80

55-2000 5-350 —

95

820 70

11

55a

17 1360

27^0 180

10.0

2 . 9 - 3 Λ

P

0.2-16.7

0.16-880

130

1070

27 1165 15^0

kko

170

66*

Isooctyl alcohol Isopropylbenzene (cumene) h,U-lsopropylidenediphenol

(bisphenol A) Lactic acid Maleic anhydride

Melamine (2 9 k,6-triamino-1>3,5-triazine)

-

3U

23 Uo -

60·

535

-90

55i

1 .2

3-2U

0.029-5 P

0 .5 -1 .0

8-61000 P

P < 5 2 . 3

P

0.01

P

P

P P

P P

P P

P

P

P P

P

P

P

P

ρ

140


Chemical

Production (10° kg approx.)

Concentration (M-g/l) in

1963(5) 1969(16") S e W a g e s u r f a c e S r o u n d t a P iyo^o; lyoiA I D ; e f f l u e n t water water water

Methanol Methyl methacryla te U-Methyl-2-pentanone

(methyl i sobu ty l ketone) H-Methyl-2-propanol ( t e r t -

1100

75 200

b u t y l a l c o h o l )

N a p h t h a l e n e 0 - N a p h t h o l N i t r o a l k a n e s (C . -C )

N i t r o b e n z e n e O l e i c a c i d P e n i c i l l i n s P e n t y l (amyl) a l c o h o l s P h e n o l P h t h a l i c a n h y d r i d e P o l y e t h y l e n e g l y c o l s

P o l y p r o p y l e n e g l y c o l s

23 16

--10

0 . 8 -

U10 215

--

1 , 2 - P r o p a n e d i o l ( p r o p y l e n e -g l y c o l )

2 - P r o p a n o l P y r i d i n e R e s o r c i n o l S a l i c y l i c a c i d S o r b i t o l S t e a r i c a c i d

S t y r e n e T e r e p h t h a l i c a c i d T e t r a c h i o r o e t h y l e n e T e t r a e t h y l l e a d

T o l u e n e

29k- & 2 , 6 - T o l u e n e d i i s o -c y a n a t e ( 8 0 : 2 0 )

1 , 1 , 1 - T r i c h l o r o e t h a n e T r i c h l o r o e t h y l e n e T r i t o l y l p h o s p h a t e Urea V a n i l l i n

680 l.k

-5Λ

Ho 25

950 -16 23

1360

57

-170 -\Λ

10U0 680

68of

--

2 5 0 d

---

7^5 320

-

75f

860

i 2d

---

2 1 1 0 d

-290

-1 5 0 0 d

I 8 0 d

1U5 270

-iUooa

-

P

P

0 . 5 - 2 5

P 0 . 0 3 - 2 0

P

< 0 . 2 5 P

2 . 1 - 1 0 0

P

3 . 9 - ^ . 2

P

8.5-9.0 8.6-9.8

20 70-UUOO

0 . 1 - 2 0 0 .001 ,

0 . 3 - 1 3 . 8 P

0 . 0 3 - 3 0

P < 0 . 2 5

P

0 . 1 3 - 8 0 100

0 . 3 2 - 1 5 0 0

90-360

3 . 3 - 3 3 0 0 0 . 8 2 - 1 3 0 0

p 0 . 0 0 1 - 1 0

P

P

0 .001 P

1 . 0

0 . 0 1 - 0 . 0 5

p 0 . 0 5 - 1 8 0

p 0 . 0 0 1 - 1 2

P p 0 . 1 - 5 5 0

P P

P

P k


Chemica l

V i n y l a c e t a t e V i n y l c h l o r i d e Xy lenes

No. of i d e n t i f i c a t i o n s F r a c t i o n (%) No. of q u a n t i t a t i v e

e s t i m a t i o n s

P r o d u c t ( 1 0 6 kg

1963(5)

185 610 190

i o n a p p r o x . )

1969(16)

3^0 170

1U50

sewage e f f l u e n t

P

18

C o n c e n t r a t

s u r f a c e w a t e r

0 . 2 - 1 0 0

U9

30

i o n ( μ β / 1 )

g round w a t e r

0 . 6

15 13

k

i n

t a p w a t e r

5 . 6 - 8 . 2 U-320

5U 1+8 26

* Can be formed by fermentation processes in the natural environment. -[* See detergents. a: 1971 b: 1962 c: 1972 d: 1970 e: 1967 f: 1968

47 42 44


APPENDIX B:

DURING

SYNTHETIC

BIOLOGICAL IN VARIOUS

Chemical

Acetaldehyde Acetic acid Acetone Acrylic acid Aniline Benzoic acid Benzyl alcohol 1-Butanol Butyraldehyde Butyric acid Catechol Citric acid Cresols Decanoic acid Ethanol 2-Ethoxyethanol Ethyl acetate Ethylene glycol Formic acid Glucoheptonic acid Gluconic acid Glycerol Isobutyraldehyde Lactic acid Laurie acid Maleic acid/maleic Oxalic acid Palmitic acid Phenol Propionic acid Salicylic acid Tartaric acid Urea

WATERS(

anhydride

ORGANIC CHEMICALS NC

SEWAGE TREATMENT(IT) AND

jTT

)RMALL^ : EASILY :

CONCENTRATIONS

Concentration

Industrial effluent

P

P

1200-16500

U-25 16000

P 8000-3330.000

50-735000 200 P P P P

500

13-7200 5-2290000

P 5-10

wu Sewage effluent

0,

0,

0, 0,

P 190 P

P

P

5 P

.03-^0 0.1 P

P

P

.1-0.3

.2-50

.03-20 70 P

20

BIODEGRADED

MEASURED

Surface waters

2-120 1.0

9.8 P

0.1-7.5 P

0.01-100

P

P

3-2U

P P P P P

0.03-30 0.1-7.0

No. of identifications 18 18 17 Fraction (%) 5^ 5^ 51


APPENDIX C: SYNTHETIC ORGANIC CHEMICALS UNLIKELY TO BE REMOVED DURING BIOLOGICAL SEWAGE TREATMENT(17) AND CONCENTRATIONS MEASURED IN VARIOUS WATERS(11)

Chemical

Aldrin* Alkylbenzenesulphonates (branched

chain) Alkylnaphthalenesulphonates Atrazine* 1,2-Benzisothiazol-3-one 1 H-Benzotriazole BHC* bis(Tributyltin) oxide Bromoxynil* t ert-Butylphenol Carboxymethylcellulose Chlorendic anhydride Chlorfenvinphos* Dichlorobenzenes Dichlorophen* 2 ,U-Dichloro-3, 5_xylenol Dieldrin* Dimethyl sulphoxide Dinoseb* Diphenyl ether Diquat* DNOC* Ethoxylated alkylphenols EDTA Hexachlorophanel" Hexamethylenetetramine 2-Mercaptobenzothiazole Paraquat* Pentac hlorophenol Pentachlorophenyl laurate Phenylmercurie acetate _o-Phenylphenol Picric acid Polychlorinated biphenyls Quaternary ammonium compounds Simazine* 2,1+,5-T*

Concentration ^g/l)

Industrial Sewage effluent effluent

P P

0.016-0.5

1-268

0.0003-6U.3

P

P

6o

25-35

11+100-5*00000 0.2-U

700 0.2-2.5 0.006-200

Surface water

0.001-1 .0

P

P

0.001-9.1 5

P

0.06-31000

P

0.001-22.5

0.19-80

60-1120 3-kQ

P

0.1-10000

P

0.00009-3200

0-0.3

No. of identifications Fraction(^) 22

7 19

16 k3

* Pesticide, common name (British Standard 1831 : 1969). t Drug, approved name (British Pharmacopoeia Commission 1970).


APPENDIX D: SYNTHETIC ORGANIC CHEMICALS WHICH SHOULD BE DEGRADED BY BIOLOGICAL SEWAGE TREATMENT AFTER A SUITABLE ACCLIMATIZATION PERIOD(1T) AND CONCENTRATIONS MEASURED IN VARIOUS WATEBS(11)

Concentration {μβ/ΐ)

Chemical

Acetonitrile Acrylamide Acrylonitrile Adipic acid Alkanesulphonat es Alkylbenzenesulphonates

(straight chain) Alkyl sulphates Allyl alcohol Anthracene Benzaldehyde Benzene Benzenesulphonic acid Biphenyl 1,3-Butadiene 2-Butanol 2-Butoxyethanol Butyl acrylate e-Caprolactam Carbon tetrachloride Chloroacetic acid Chloroanilines Chlorobenzene Chlorobenzoic acids Chlorocresols Chloroform Chlorophenols 3-Chloropropene Chlorthiamid* Cyanuric acid Cyclohexane Cyclohexanol Cyclohexanone Cyclohexylamine 2,U-D* Dalapon* 2,1+-DB* Dihutyl phthalate Dichlobenil* 1 ,2-Dichloroethane 2,U-Dichlorophenol 1 ,2-Dichloropropane Dichlorprop* Dichlorvos* Diethanolamine Diethylamine Diethylene glycol Di(2-Ethylhexyl)phthalat e

Industrial effluent

0.2-U2 100000 3700

1.6-7.0

2.5

30

250

P

10

500

Sevage effluent

280

700-12500

P P P

1-80

P

P

P 0.26-1 .1

P 0.51-1 .7

P

55-250

Surface water

0.3-1 .2

P P P

10-100000

1 .0

0.16-200

O.OU-1

0.3-1700

1.5-2.2 0.16-170

P P P

P 10 1 .0

0.01-0.91

5-350

P 0.16-880

6.6 0.0U

10

10


Chemical

D imet hylami n e Dimethyldithiocarbamic acid D imet hy If ormami de Diuron* 1,2-Epoxypropane Ethanolamine Ethoxylated aliphatic acids Ethoxylated aliphatic alcohols Ethoxylated alkyl amides Ethoxylated alkylamines Ethoxylated alkyl sulphates Ethyl acrylate Ethylamine Ethylbenzene Ethylenediamine 2-Ethyl-1 -hexanol Fluoroac etamide Fluoroacetic acid Formaldehyde Formamide Fumaric acid Glycollic acid Hexane k-Hydroxybenzenesulphonic acid Malathion* Maleic hydrazide MCPA* Mecoprop* Methanol 2-Methoxyethanol Methyl acrylate Methylamine Methylene bisthiocyanate Methylene chloride Methyl methacrylate 1-Methylnaphthalene 2-Methyl-2,U-pentanediol U-Methyl-2-pentanone 2-Methyl-2-propanol Morpholine Naphthalene Nitrilotriacetic acid Nitrobenzene Nitrophenols Nitrotoluenes Octadecylamine

Industrial effluent

P

P

P

19

1000

27

3-100

53-1000

110

1 5-7800

Concentration L

Sewage effluent

P

P

1 .0

P

2.9-3Λ

1.0

P 170-1560

1.U 12

g/1)

Surface water

P

P 1 .2

U1

P

0.3

0.3 0.0U-1.3

P

0.6-Uo

0.1-20

0.3-13.8 1.3

0.25-20


Concentration ^g/l)

Chemical Industrial effluent

Sewage effluent

Surface water

1 -Octanol Oleic acid Paraformaldehyde Parathion* Pentaerythritol Phenanthrene Phenylacetic acid £-Ph eny1enediamine Phlorogluc inol Phthalic acid/phthalic anhydride Picloram* Polyethylene glycols Polypropylene glycols Polyvinyl acetate/polyvinyl alcohol 1 ,2-Propanediol 2-Propanol Pyrethrum Pyridine Pyrogallol Quinol Resorcinol Rotenone* Stearic acid Styrene Sulphanilic acid 1,1,2,2-Tetrachloroethane Tetrachloroethylene Tetrahydrofuran Thiourea Thiram* Toluene Toluene-2 5U-diisocyanate £-Toluenesulphonic acid Toluidines 1 ,1 ,1-Trichloroethane Trichloroethylene Trichlorophenols 1,1,2-Trichloro-1,2,2-trifluoro-ethane

Triethanolamine Triethylamine Triethylene glycol Trimethy1amine Vinyl acetate Xylenes Xylenols

No. of identifications Fraction {%)

19000 80-U6000

70-1UOOO

P

5000-23UOO

P 150000

20.200

2.6-31

65Ο

10-10000

200 25-115

P

30 820-13Ö000

36 26

0.5-25

P P

P

<0.25 <0.25

2.1-100

P

3.9-^.2

P

P

8.5-9.0 8.6-9.8

P

P P

kO 29

P

0.005-0.065

0-1 .3

P P P

P

<0.25

<0.25

P 0.13-80

2.3-6UO 0.32-1500

90-360

0.7-2 3.3-3300 0.82-1300 0.05-0.2

P

P

0.2 P

61 UU

* See footnote to Appendix C.

Pollution Abatement; An Industry's Point of View N. VAN LOOKEREN CAMPAGNE

SEBB = Foundation ofEuropoort Industries, Rotterdam

The subject of this presentation for the 2nd International Symposium on Aquatic Pollutants is "an industry's point of view" or better "the point of view of those Rotterdam industries" which work together on environmental issues in an organi-zation called the "Europoort Botlek Industrial Interests".

The presentation is roughly divided into four parts:

Commencing with some introductory remarks on tackling environmental problems in general, the field is then narrowed to the transnational problem of cleaning the Rhine, as an example of water pollution abatement and aquatic pollutants.

The "analytical part" that follows is the main item, advocating registration and publication of all emissions entering the Rhine basin as an integral part of an allround scientific/technocratic approach - you may call it a philosophy -of tackling water pollution abatement. In some important aspects this part is closely connected with Dr. van Esch's paper presented on the 26th.

The next part deals briefly with some "government and administrative aspects" of cleaning our surface waters, in this case the Rhine. This subject, however, is dealt with in greater detail by Mr. de Geer's talk to be given on the 28th on "Regulatory actions in Europe relating to aquatic pollutants".

The final part is mainly a plea for closer co-operation between the scientific institutions, politicians, industries and civil servants of the countries involved and the deployment of all expertise available to clean our surface waters.

INTRODUCTORY REMARKS

The speaker is a Shell employee, responsible for co-ordinating environmental affairs for the ten local Shell operating companies inside The Netherlands; in ad-dition to this work, however, he also participates on behalf of the oil industry in the councils, committees and public institutions in Holland, in which a solution to the environmental problems is jointly sought.

169

170 N. van Lookeren Campagne

In this context, one of the speaker's jobs is in the Stichting Europoort Botlek Belangen, the foundation representing the interests of 60 major industries located in the 20 miles between Rotterdam city and the North Sea. This industrial complex belongs to the largest in Europe, it is closely connected with the largest port in the world and situated in a rather populated area of lj million people living in Rijnmond (a kind of a Greater Rotterdam), which is roughly 20 by 30 miles and located at the end of the Rhine. It is with this back-ground that this presentation is given.

The Rotterdam area of course has its share of environmental problems: air and water pollution, waste disposal problems, noise abatement, etc. If you add to this the important Dutch aspect of "Land Use Planning" (the land of which we have precious little), the safety aspects within and around the industries, energy conservation, recycling of materials etc., then you have the "environmental issues" in a rather great variety. From the work in this Foundation, many of us (about fifty, most with a university background, working quite close together), have learned two lessons:

The first is an analogy to the first Law in thermodynamics on the conservation of energy and in jokes it is called the "Law of conservation of environmental misery". If you clean up effluents, beware of the waste-disposal and smell problem; if you solve this by incineration, beware of air pollution; if you switch from water cooling to air cooling in order to decrease the thermal burden on the surface water, you create a noise problem, etc.! Sometimes authorities dealing with a certain pollution problem or scientists investigating a specific environmental issue, tend to forget this "Law on conservation of environmental misery"; in industry, however, one is often confronted with the integrated issue. Generally, we deal in industry with the problems of improving existing plants or processes, but the best way out of this misery-dilemma is of course to work on a new generation of processes in which from the conception stage onwards, environment aspects have been included.

The second lesson is: Don't see pollution problems as a semi-static process in which emissions just continue to fill up our precious environment until it bursts; but treat it as a dynamic process of "sources" (the emissions, the effluents entering the river), the spreading of the emitted components in the surrounding environment in which it may change physically or chemically, and the "sinks" (as opposed to sources) in which it leaves the environment as a specific polluting component. In this dynamic process there is often some sort of stability and we will find the polluting component at a certain steady-state level, a level which may fluctuate due to human or natural disturbances. Human disturbances are for instance incidents like an oil spill, natural disturbances are for instance the temperature fluctuations of the seasons, the increased waterflow in spring with an increased amount of inorganic components, or the increased charge of organic material in autumn. So, air pollution in Western Europe, cleaning up the Rhine, solving waste disposal problems in Rotterdam or finding a noise abatement stra-tegy around an industrial complex: they all can be roughly described in terms of sources, spreading and changes of components, sinks and a resultant pseudo-steady-state level of the component in this dynamic process. This theme will be worked out in my presentation.

Narrowing now the field to water pollution abatement and aquatic pollutants, the Rhine problem is taken as example for the rest of this presentation. Not only the river from Basle to Rotterdam, but the Rhine basin with all its tributaries, ending in the Rhine estuary in Holland, influencing our coastal waters of the North Sea, influencing also the inland sea called IJsselmeer and the large shallow tidal waters 100 miles north of the mouth of the Rhine, called Waddenzee. Although this talk is now focussed on this Rhine problem, the apprach is thought to be quite generally applicable.

Pollution Abatement 17ί

ANALYTICAL PART

We now come to the analytical part of the presentation. What is still lacking, even after much work has been done, is a proper registration of all effluents entering the Rhine basin. The nearest emission survey is the report "Umweltpro-bleme des Rheins" published in 1976. I am quite aware of many arguments against an all-embracing emission registration and selective publication of its results, directed at improvement of the surface waters. If I advocate here such a thing -taking into account besides effluents from industries also the contributions from towns, villages, shipping, nature etc. - I do so on the basis of experiences with a general emission registration system in The Netherlands, which started about 4 years ago and will take another three years to cover the whole country.

This Dutch Emission Registration system has the following aspects:

It is initiated, supervised and financed by the Ministry of Public Health and the Environment; it is mainly carried out by a well-known R & D institution in Holland, called TNO. All emissions to air and water and the chemical wastes from industry are regis-tered.

Essentially, it is the same system as applied by TUV in some parts of Germany, where they started around 1971 in the Cologne area. It is quite an elaborate system, generating an enormous amount of data in a computerized data bank, with hundreds of air- and water-polluting components and many details on how, why, when and in what quantity these emissions occur.

Recently an initial report has been published, dealing with an emission survey of part of the province of Limburg. For the present presentation it is sufficient to state that a vast amount of ex-perience exists on how to get and administrate emission data and some experience on how to make these data available in surveys which give a clear picture of what is really going on in our environment, so that authorities and other parties con-cerned can plan their pollution abatement measures in an optimal way.

Such a system would probably be quite sufficient for the registration of all rele-vant components covering all important effluents entering the Rhine basin. If anything, the system should be simplified, but for the rest a lot of teething troubles have been dealt with in Holland and Germany. If all parties concerned with the Rhine effluents really agree to start emission registration, then the hard and soft ware is ready and waiting. It is clearly more a question of poli-tical willingness then technical impossibilities.

Here some information has to be added which logically should come under the next part "government aspects":

There is no law on emission registration in Holland; the whole thing is carried out on the basis of free co-operation between the central Government, the pro-vincial and local authorities, a public institution and last but not least industry. There are firm agreements between the parties concerned, but in essence the co-operation is based on mutual trust and a general public spirit to make a combined effort to collect sufficient data to find the best measures to combate pollution. Industry has from the very beginning been in favour of full co-operation, and the Dutch Employers' Association together with the Chambers of Commerce in Holland have appointed two industrial representatives to the Emission Registration Board which is headed by the Director General of the Environmental Department of the Ministry of Public Health and Environment. There are plans to issue in due course a bill on emission registration when enough experience has been gained to see our way clearly in this complicated and sometimes quite delicate field.


Delicate, because for industry - and other parties as well - it is very important to know from the very beginning what "the rules of the game" are and also to know that all parties will adhere to these rules. In Holland these rules are:

First there shall be a rough overall emission registration, then a more detail-ed registration will be made. The registrations will include all emissions (also from nature, transfrontier pollution etc.) and work on all these emissions will be carried out at the same time (not industry first and the rest later).

The effort will be geared to a complete picture of emissions, the spreading-out of components, physical/chemical changes, resultant immissions etc. (i.e. not limited to emissions only).

Publications of the results will be focussed on a good understanding of environ-mental pollution and not on publishing individual industrial emissions (of course big industrial complexes which dominate their environment can be recog-nized from the published data, but that is consistent with the general philoso-phy of understanding the polluting phenomena and has nothing to do with identi-fying individual polluters).

Only those authorities appointed by law to implement certain pollution abatement measures have access to all data within the sphere of their operations.

We feel that Dutch industry has taken a wise decision in co-operating from the very beginning and we think that industry is entitled to proper care of their spe-cific industrial interests by the authorities. Thus, our mutual interest is that the best (longer term) pollution abatement programme will be one based on the best data available, coupled with a good abatement philosophy. This brings us back to the second section of the analytical part: "studies on what happens with emissions". What is needed for the Rhine clean-up is not only regis-tration of all effluents (including natural emissions) and a proper survey of all emissions etc. Merely to curb emissions, regardless of what their effects are, should not be our aim. Some people do not look further ahead and just say: "all emissions should be zero, if not zero then in any case as low as is possible with all available technology". This, however, seems to be a rather narrow point of view. Industry's point of view is that we have to have the whole picture and this view has been adopted in the Dutch Emission Registration Board, although by some with more enthusiasm than others (who probably think of the administrative simplicity of just curbing emissions as compared with the more complex public administration if decisions are based on "the complete picture").

The third section of the analytical part of my presentation advocates three items:

A registration of the use to which the waters of the Rhine are put at present.

A good understanding of the assimilative capacity of the various receiving waters at the moment.

and a clear view of the improvements and changes that we want to obtain in the future with regard to the first two items.

Closely related to the registration of the "use to which the waters are put", is the work on quality standards for each of these types of water.

On all these aspects there are data and there is some experience; probably the least is known on the aspect of assimilative capacity. But it is also clear that on all the topics just mentioned far more detailed and isolated know-how is availa-ble than is currently published in a systematic and general view, focussed on a proper cleaning operation of the Rhine. Everywhere we find isolated bits of information, but as far as I know there is no

Pollution Abatement J73

published systematic survey of the Rhine basin (except the above-mentioned "Umwelt-probleme des Rheins" which is a first step in the right direction). If one looks at the large interests and staggering amount of money concerned, it must be a great challenge to tackle the job of systematically compiling the available know-how, to point out the gaps in know-how and to distribute the work between the scientific institutions in Europe and abroad in order to fill these gaps. But this subject already belongs to the next part of the presentation.

GOVERNMENT AND ADMINISTRATIVE ASPECTS

How do we get harmonized emission registration laws (or agreements) in the various countries involved?

How do we get registration of the use to which the waters are put and the assimilative capacity of the receiving waters?

How do we get proper agreements on the use of registered data and multinational co-ordination of the various studies indicated earlier?

How do we work out harmonized water quality standards for the various purposes to which we intend to use the water?

These questions are raised here, but not answered. This presentation deliberately makes the somewhat artificial division between the "analytical part" and "govern-ment aspects". It is probably good to see as clearly as possible in such a com-plicated problem as cleaning up the Rhine, what the scientific, technical and eco-nomic contributions to a solution are or can be, and what belongs to the sphere of general governments organization: international co-operation, co-ordination, diplomacy, trust etc.

We refrain from making a choice between the various possibilities to solve the questions such as:

A supranational organization. An example of it is the Central Rhine Shipping Committee, which has functioned since 1815 and regulates free shipping on the navigable part of the Rhine. The regulations are compulsory, some of them have an environmental aspect such as spills from ships; infringements are dealt with in special courts, there is a special court of appeal.

Faure's idea of a "Bassin Rhenane", along the line of the French "Agences de bassin" and more or less the Dutch Waterschappen and the recently created English Water Authorities.

- The International Commission for the Protection of the Rhine.

Initiatives from the European Communities to solve the Rhine problem.

Initiatives from the Parliaments of the various countries, working together in preparing legislation etc.

There is a great variety in possible structures to manage and govern the clean-up of the Rhine; Mr. de Geer's paper on the 28th deals with this in greater detail. As far as is known all of them have been tried or proposed with only limited suc-cess and according to realists only a limited scope for success in the near future. Industrial organizations have hardly been involved in these endeavours. Many people in industry, however, feel that we must somehow reach an agreement on cleaning operations in which the various waters are restored to a state in which they will fulfil better the function or functions we want them to perform. These functions can be:

Preparation of drinking water. Water to breed fishes for consumption.


Water with a high assimilative capacity or an optimal eco-system. Drinking water for animals. Water for various agricultural uses. Water for recreation, swimming etc. Water for industrial use, process water, cooling water etc. Water just for shipping. End-of-river water which is of a quality that it does not adversely affect the coastal waters and estuaries with their specific functions etc.

Each function roughly required certain water quality standards, and these standards together with the properties of the receiving water will determine the requirements for the effluents. Of course, salt emissions at Strasbourg are quite different in their effect there, compared with the same emissions in the salt tidal waters at Rotterdam; on the other hand, requirements regarding sea-water eco-systems might be more stringent in Rotterdam than for instance in an inland tributary passing through a major industrial area.

At the moment, quite a number of effluent-treating projects along the Rhine have been carried out or are under way, determining the work and expenditure for the first few years. To pick the first obvious cases is not so difficult; the diffi-culty will come when the law of diminishing returns (less and less environmental benefit for more and more money) begins to show. What we will then increasingly need - besides good data and a sound basic philosophy as stated earlier - are careful water-improvement schedules taking into account:

timing costs; financing of projects; various economic aspects all kind of cost/benefit considerations socio-economic consequences

(On purpose industrial competition is not listed here as a factor).

FINAL PART

As the time available for this presentation has been curtailed, we will skip some light criticism from the EBB Foundation on - what we in industry regard as - a lack of communication with some groups of civil servants and some scientific insti-tutions, and on an insufficient use of industrial know-how in solving pollution problems. The Foundation and its industries advocate closer co-operation between the scientific institutions, politicians, industry and civil servants of the countries involved, as well as international bodies such as the EEC. We feel that these multi-multi-billion (10 ) guilder clean-up operations on our surface waters require more extensive studies and co-operation than they receive at present and the deployment of all expertise available.

Summing up

This presentation has attempted to give a rough idea of an industry's point of view on pollution abatement in general and on cleaning-up the Rhine in particular. Most of it is not really new. It is obvious that some of the ideas presented here still need brushing up and the help of scientific institutions to solve the problems involved would be welcome. Maybe one aspect is new: industry being in favour of a general Rhine effluent registration on condition that the results are used in the ways as indicated in this paper. Maybe this viewpoint - based on positive experiences with emission registration in The Netherlands - will help us somehow to get out of the blind alley, the deadlock, which we seem to be in at present with the cleaning-up of the Rhine. And it may be that the underlying simple philosophy will help to find a way in improving the surface waters in general.

Volatilization of Pollutants from Water

DONALD MACKAY

Department of Chemical Engineering & Applied Chemistry and Institute for Environmental Studies, University of Toronto, Toronto, OntarioM5S 1A4, Canada

ABSTRACT

The mechanism of the process of volatilization from water is briefly reviewed in terms of liquid and vapour phase resistances coupled by the Whitman Two Film Model. It is shown that volatilization rates are principally controlled by three variables, the Henry's Law constant, the vapour phase mass transfer coefficient, and the liquid phase mass transfer coefficient (which is related to the reaeration constant commonly used in studies of oxygen transfer into water). The current state of knowledge of values of these three variables is reviewed, and methods of experi-mental determination and estimation are outlined. The effects of adsorption and surface layer phenomena in modifying volatilization rates are discussed briefly.

INTRODUCTION

Aquatic pollutants which do not ionize in aqueous solution often volatilize, occasionally at high rates, leading to surprisingly short residence times in the lake, river or ocean. It is important to develop methods of calculating volatiliza-tion rates for such compounds if their environmental behaviour and effects are to be understood, and controlled, and ultimately modelled mathematically. Volatiliza-tion rates depend on a number of thermodynamic or physical-chemical properties of the pollutant, especially aqueous solubility, vapour pressure,diffusivity and Henry's Law Constant which in many cases are not accurately known because of the difficulty of experimental measurement. The rates also depend on the fluid mechanical regimes which exist in the water body and in the lower atmosphere. Unfortunately, these regimes change from hour to hour; thus for a given location the volatilization rates may change by a factor of ten over a period of a few hours, thus rendering difficult, if not invalid, the use of simple average rates. A further complication is that volatilization rates of a pollutant may be profoundly influenced by the presence of other substances, notably adsorbants, organic materials which accumulate at the air-water interface, and electrolytes.

175

176 D. Mackay

In this review, an attempt is made to discuss these processes and how they may be quantified in calculating volatilization rates. Some gaps in knowledge are identified and means of obtaining the required information suggested. It is emphasised that successful calculation or correlation of volatilization rates can only be achieved if the equations used are reasonably representative of the prevailing physical situation; thus it is appropriate to examine first the mechanism of transport of a pollutant from the bulk of a water body to the atmosphere.

MECHANISM OF VOLATILIZATION

Ideally one would like to be able to observe the microscopic behaviour of the pollutant molecule as it moves from the water column, to and through the surface, and out into the atmosphere. Only indirect evidence of the nature of this two phase diffusion process is available, principally from measurements of concentra-tion profiles. Thus there are still uncertainties and this is reflected in an inadequacy of correlation and prediction about some processes, particularly at an interface. It is convenient here to discuss separately the processes in the bulk of the water and air phases and in the immediate vicinity (i.e. within a few millimetres or centimetres) of the interface.

Interfacial Processes

It is generally accepted that most of the resistance to volatilization lies in the few millimetres or centimetres above or below the interface. These mass transfer processes are of industrial importance in gas absorption, gas stripping and distillation and it was the pioneering work of Whitman in the 1920's which established the first useful model in which the two phase resistances were separated and the equilibrium was assumed at the interface. The two phase resistances, usually characterised by a mass transfer coefficient, can then be treated separately. The Whitman model has been successfully used by Liss (1,2), Liss and Slater (3) and Mackay and Leinonen (4) and others to calculate transfer rates. The simple mathematical relationships are presented below.

2 The mass flux N (mol/m h) across the interface is expressed as a function of the liquid phase concentration C(mol/m ) , the vapour phase partial pressure P(atm),

3 the Henry's Law Constant H (atm m /mol) and the overall liquid phase mass transfer coefficient IC (m/h). This coefficient can be further expressed as a function of

the two individual phase mass transfer coefficients, k for the liquid and kn for -5 3

the vapour (both m/h), the gas constant R (8.2x10 m atm/mol K) and the

absolute temperature T (K).

N = K^C - P/H)

Ι/ί^ = 1/1^ + l/(HkG/RT)

The mass transfer coefficients can be regarded as conductivities in an Ohm's Law sense and their reciprocals as resistances, thus the latter equation is essentially the addition of two phase resistances in series to yield an overall resistance. It is thus possible to calculate the relative contribution of each resistance and often one resistance dominates. If k_ << (Hk /RT) the liquid phase resistance

controls and the flux equation becomes

Volatilization of Pollutants from Water 177

N = kL(C - P/H)

For most, but not all, hydrophobic pollutants the liquid phase resistance dominates and the atmospheric concentration is negligible thus the transfer rate is independent of H. In general, however, it is essential to have values for all three quantities; k_, k and H. Environmental values of K and le. are such that

the ratio k /k^ normally lies in the range 50 to 200, Liss (3) suggesting a value for oceans about 96. The percentage resistance in each phase can be expressed as

a function of H, as shown by Sutherland (5) in Fig. 1. For H values below 2x10 , the pollutant tends to partition into the liquid (i.e. it is relatively soluble)

-3 and the transfer is gas phase controlled. Above 10 , the liquid phase controls. Interestingly, many polynuclear aromatics and halogenated aromatics of environ-mental concern lie in the intermediate range where both resistances are important.

To elucidate the nature of the physical processes at the air-water interface,the ideal experimental approach would be to measure concentrations in the interfacial region, especially in the 1 mm below the interface. This is presently impossible. Some information can, however, be obtained on profiles in this region if one invokes the analogy between heat and mass transfer and assumes that turbulent(or eddy) heat and mass diffusivities are equal. The near-interface region can be examined on a microscopic scale by plunging a small thermocouple or thermistor through the interface and measuring the temperatures. This technique, originally used by Holley (6), has been recently repeated by Pochmursky (7) who has obtained temperature profiles using a thermocouple about 0.02 cm diameter with a very fast response. A typical profile illustrated in Fig. 2 shows that most of the tempera-ture gradient lies within one mm of the interface. Knowing the temperature gradient and the heat flux, the total thermal diffusivity can be calculated. The molecular thermal diffusivity can then be subtracted to leave the eddy diffusivity. As expected, this tends to zero at the interface where it is impossible to have a vertical velocity at a horizontal surface. From microscopic or radiometric measurements, it is hoped that the fundamental mechanism of interfacial transfer can be elucidated and thus predictive equations devised which have a sound basis in physical reality. In essence, the problem is to understand how turbulence generated in the atmosphere or in the water changes as it penetrates the interface where, in theory at least, surface forces constrain or eliminate the vertical movement of eddies. An example of the need for greater fundamental knowledge is the speculation that the extent of water evaporation, direct heat transfer or solar radiation absorption at the water surface may profoundly affect mass transfer. If the water surface is cooler than the bulk, buoyancy will tend to assist sinking of denser surface elements. Conversely when the surface is warmer, stability and low transfer rates may ensue. A complication which is discussed later is that the interfacial region may be contaminated by floating or surface active material which modify interfacial processes, generally damping turbulence and providing a resistance to mass transfer.

Bulk Water Phase Processes

The vertical transport of pollutants in a water body is controlled by currents (both direct water currents as in rivers or estuaries and wind-induced currents) and the presence of stable regions such as thermoclines. Generally horizontal turbulent diffusion is an order of magnitude faster than vertical turbulent diffusion suggesting that horizontal layers of water slide over each other, only occasionally mixing vertically. The "roll cells" can thus be conceived of as horizontally long and vertically thin. It is impossible to generalise about such processes since there are so many different hydrodynamic configurations but it is worth examining briefly as examples the different characteristics of two regimes,

D. Mackay

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rivers and lakes.

In rivers, vertical transport ;Ls fast and is dominated by eddies caused by the interaction of the current with the river bottom. The extensive studies of river reaeration have shown that the resistance lies at the surface. Normally the reaeration constant K« has a value of about 1.0 (day" ) for a river of depth 2 m

flowing at about 1 m/s (8). The mass transfer coefficient V. and the reaeration

constant K« can be shown to be related by the equation k =DK« where D is the

river depth. Thus a K of 1.0 day with H of 2 m corresponds to L of 2 m/day

or 8.3 cm/h. This value is comparable to those suggested by Liss (3) for transfer at ocean surfaces. The usual correlation for K? takes the form

K = (constant)Vn/Dm

where V is the stream velocity. The power m is usually 1.3 to 1.7,thus the mass transfer coefficient is inversely proportional to H to the power (m-1) i.e. 0.3 to 0.7. A physical interpretation is that turbulence is generated at the river bottom and that the more distant the bottom from the surface (where the resistance lies) the more the turbulence is damped and thus the more quiescent is the water surface and the greater the resistance to mass transfer.

In contrast, in lakes without significant flows, the turbulence originates in the atmosphere, thus windspeed is believed to be the controlling parameter. As a result of wind stress on the water surface, waves are generated which induce turbulence beneath the surface. The near surface layer exhibits very low vertical diffu-sivities but at depths of a few centimetres to several metres vertical eddy diffusivities can have values of about 30 cm2/s or six orders of magnitude greater than molecular diffusivities (9). The diffusivity then tends to drop to low values at regions of thermal stability (thermoclines). Calculating the vertical transport of a pollutant through these regions of differing diffusivity is thus complex, however, since concentration gradients tend to be inversely proportional to diffusivity for a steady mass flux (Fick's Law) the regions of greatest resistance (and thus greatest interest) are thermoclines and the interface. Emerson (10) has published examples of concentration and temperature profiles in a small lake and Broecker and Peng (11) have published similar oceanic data. Any real understanding of the volatilization process (and ultimately its modelling) for a given water body must be based on a detailed knowledge of the prevailing hydraulic regime, particularly the vertical eddy diffusivity as a function of depth and time.

Another important process is the exchange of pollutant between bottom sediments and the water column. The sediments may act as a temporary or permanent sink for pollutants. The elucidation of this process is a fascinating and difficult problem involving geochemical, adsorptive, fluid mechanical, biological and microbiological considerations. Our ignorance of the potential of these processes to accumulate large quantities of toxic persistent materials such as mercury or PCBs then release them back into the water column has been responsible for severe environ-mental problems in the Great Lakes.

Bulk Air Phase Processes

Normally vertical diffusion in the atmosphere is sufficiently fast that a vola-tilizing pollutant will be transported from the near-surface very quickly. Only under extremely stable atmospheric conditions of temperature inversion do the rates become small and even then the resultant calmness of the water causes an

180 D. Mackay

increase in water phase resistance. Bulk air phase processes are thus normally quantified when the aim is to construct a plume dispersion model or when a very soluble pollutant such as SO« is being absorbed by the water. For the present purposes they can be ignored.

THE LIQUID PHASE MASS TRANSFER COEFFICIENT le

A considerable volume of information is accumulating on values for k_ in both environmental and laboratory systems. For rivers, the many oxygen transfer studies have provided correlations for the reaeration constant (and hence L ) as a function of stream velocity and river depth. In lakes and oceans there have t>een studies of transfer of several gases but notably radiocarbon C09 and Rn by Broecker, Peng

and Emerson and Thurber (10,11,12,13) . Windspeed is the dominant variable and several correlations have been proposed notably of the form log k as a function

of windspeed, and L as a function of windspeed squared. Rather than review this

fluid mechanical problem in detail, it is possibly of more interest to describe some studies by Cohen (15) which throw some light on this topic.

Cohen measured k-r over a range of windspeeds from 0 to 12 m/s in a tank 2.4 m long by 0.62 m wide by following the decay in concentration of dissolved benzene and toluene. Provision was made to stir the tank contents, but not vigorously enough to deform the surface. Velocity profiles were measured in the air phase and correlated to give the surface roughness Z (an approximate measure of the

waviness) and the friction velocity U (a measure of the stress of the wind on the water), as well as the free stream velocity U^.

The results shown in Fig. 3 can be interpreted as follows. Below 3 m/s, le. has values below 3 cm/h with relatively little effect of windspeed. The dominating variable is stirrer speed or turbulence induced from within the water. Apparently the wind stress is insufficient to transmit much turbulence into the water. This is in agreement with Emerson (10) who reached a similar conclusion for the R data in lakes. Probably under these calm conditions the volatilization rates are controlled by water currents or residual turbulence from previous winds. Such calm periods will result in low volatilization rates and possibly the accumulation of volatilizing pollutants in the near-surface waters.

In the range 3 to 10 m/s, which is probably of greatest environmental significance, k increases substantially to about 30 cm/h due to the onset of rippling and waves. Cohen successfully correlated k with U* and Z suggesting that wind induced

roughness and stress controls mass transfer.

A third region is likely to occur above 10 m/s where wave breaking may commence. It is difficult to obtain either laboratory or environmental le. data in this region.

Since kT is not a linear function of wind speed and probably also depends on fetch and water currents, it is difficult, if not impossible, to calculate a mean value. At times of high volatilization the surface waters may become depleted of the pollutant and the rate may then drop because of the drop in the concentration. Calculations of annual amounts of a pollutant volatilized using an average k and and average concentration could be considerably in error because of such effects. A tentative deduction from Fig. 3 is that the contributions to k from wind and stirring are additive, suggesting that windspeed and reaeration-type correlations could be combined for water bodies where current and wind are important in generating turbulence.


It is clearly impractical to measure k values for all individual pollutants. Studies of turbulent transfer suggest that the process is principally movement of the pollutant contained in an eddy or fluid element. Although this eddy "velocity" will be of molecular properties, experiments show some dependence on molecular diffusivity. This can be explained in terms of a period of unsteady state molecular diffusion (characterised by Fick's Second Law) occuring when a fluid element is exposed to the interface. Most mass transfer studies suggest a dependence of le on molecular diffusivity raised to a power between 0.5 and 1.0, usually about 0.67 (16). Since correlations for diffusivity as a function of pollutant molar volume are available, this provides a rational method of calculating k~ for one compound from values for another. Liss (3) has suggested using the square root of molecular weight to correct k_ and Tsivogolou has used molecular diameter (17). This correction is still not as well established as is required, particularly if k_ values for low molecular weight molecules such as oxygen are to be used to estimate values for compounds such as PCBs.

THE GAS PHASE MASS TRANSFER COEFFICIENT kn

Fortunately the extensive micrometeorological work on evaporation from lakes has resulted in an adequate knowledge of k as a function of atmospheric conditions. Values are typically about 3 m/h. If a logarithmic velocity profile is found or assumed, k can be calculated from measurements of concentration made at two vertical heights. This has been discussed by Liss for SO« absorption and water evaporation (2) and an example of SO« transfer measurement has been given by Whelpdale and Shaw (18).

HENRY?S LAW CONSTANT H

An accurate value for H is essential for any estimate of volatilization rate. In some cases where the pollutant concentration is measurable in both air and water, the value determines the direction of transfer. Usually it is regarded as the factor determining the distribution of resistances and as such it can profoundly affect the transfer rate. When H is small, (the gas phase controls) and the pollu-tant partial pressure in the atmosphere is negligible, the volatilization rate is proportional to both H and k .

H is conventionally determined by measuring a gas partial pressure and a liquid concentration. Greatest accuracy is obviously obtained at high pressures or con-centrations, usually orders of magnitude above environmental values. This is thus a risk that H at environmental conditions may differ from the "Handbook" value. Although some results for CO suggest that this may be the case (2), there are thermodynamic difficulties in explaining such behaviour at high dilution. For this and other reasons, for example the possible modification of H by low concentrations of dissolved or suspended material such as electrolytes, fulvic or humic acids or minerals, it is clearly preferable to measure H directly at low concentrations using environmental water. This is particularly important for high molecular weight compounds such as polynuclear aromatic hydrocarbons, PCBs and some pesti-cides which have very low vapour pressures and low solubilities, but are highly hydrophobic. Their hydrophobic nature causes them to exert a much higher vapour pressure than would be calculated from Raoultfs Law, i.e. they have high activity coefficients. Accordingly, they are often rapidly volatilized from aqueous solution. For example, naphthalene with a vapour pressure of only 0.23 mm Hg at 25°C will volatilize from a i m deep volume of water with a half life of about 7 hours (4). The ratio of naphthalene to water in the vapour is very much higher than in the liquid, i.e. the relative volatility is high despite the low

182 D. Mackay

naphthalene vapour pressure.

Experimental determination of H for such compounds is thus difficult and often inaccurate because errors are introduced during measurement of two small quantities, a vapour pressure (which may have to be extrapolated from higher temperatures) and a solubility. A more promising approach recently devised by Sutherland (5) is to follow the concentration decay in a volume of water which is being stripped of the pollutant. It can be shown that the slope of the log concentration versus time line gives H directly and as the concentration falls the dependence of H on concentration can be observed. Both the kinetics and equilibria of added adsorbing species or electrolytes can be investigated directly.

There is a real possibility that predictions of volatilization rates based on "distilled water" physical chemical data may be substantially too high when applied to low concentrations of pollutants from "environmental water" because of the adsorptive effect of natural organic or mineral materials. It is easy to conclude that a substantial fraction (90% or 99%) of a pollutant present at a con-centration of 1 mg/m-* is adsorbed on natural adsorptive material present at a typi-cal level of 1000 mg/m . Such an effect could reduce volatilization rates by one or more orders of magnitude since only a small fraction of the pollutant would be truly in solution and thus capable of exerting the vapour pressure or fugacity which drives the volatilization process. The lack of quantitative data on this effect, especially at low environmental concentrations represents the major gap in present knowledge and predictive ability.

It should be noted that this discussion applies to non-reacting or ionizing pollutant species. When reactions occur, as with CO« or S09, carboxylic or

phenolic acids, it is necessary to take the reaction equilibria into account and calculate the concentration of the volatilizing species in the water, rather than the total concentration of all derived species including ions.

Since measurement of H is difficult, there is an incentive to develop methods of calculating or predicting H from easily accessible basic pollutant properties such as molecular weight, molar volume, melting or boiling point and molecular structure. For pollutants which are solid at environmental temperatures (eg. many pesticides and polynuclear aromatic hydrocarbons), it can be shown that H is dependent on the solid vaDour pressure, the vapour pressure of the subcooled liquid and the activity coefficient (or excess Gibbs free energy) in aqueous solution. All three are amenable to correlation with readily accessible properties and considerable progress has already been made,for example, using a group contribution approach. Predictive methods of this type are invaluable not only for obtaining data for compounds for which no data exist but also for testing existing data to reveal discrepancies which arise from faulty experimental measurements.

SURFACE FILMS

It is well established that surface films, usually believed to consist of surface active organic material, impede transfer processes at an interface. This may be a "blocking" effect or due to damping of near-interface turbulence. In heavily polluted waters where visible scums or slicks are present, there is little doubt that volatilization rates will be depressed. This topic, which has been reviewed recently by Liss (19) is complicated by difficulties in sampling, observing and analysing these films and by their probable site-specificity. The extent to which they reduce volatilization rates is not well understood.


MODELLING

Models of aquatic processes can have several purposes; to elucidate interactions between physical and biological processes; to provide verification that the principal processes in an aquatic system are sufficiently understood; and for regulatory purposes to test the effect of control measures or new pollutant inputs. Incorporating volatilization rates into such models is probably one of the simpler components, although there may be doubt about the values of certain parameters. One of the most difficult aspects may be determining the time intervals which should be used in assembling numerical models. As was discussed earlier, there may be profound diurnal effects, especially in small lakes, when the pollutant is largely depleted from the surface waters during times of high turbulence, with a build up again during quiescent periods, for example at night. There is little doubt that effective models of aquatic systems are destined to be very complex, involving a large array of interacting physical, chemical and biological processes. The formidable task faced by those who assemble such models and attempt to provide numerical solutions can be helped if the volatilization process is described in simple conceptual and mathematical terms. This is particularly important in the first stages of modelling where simple compartment models may be used with simple rate expressions or half lives (4).

CONCLUSIONS

The development of a full understanding of the volatilization process and ultimately the prediction of rates for any pollutant in any water body under any hydrological or meteorological conditions presents a challenge to many of the traditional disciplines which contribute to environmental science. There are problems in meteorology, hydrology, fluid mechanics, physical chemistry and thermo-dynamics, interfacial phenomena, analytical chemistry and adsorption, in exploring and quantifying the interaction of volatilization with other processes, both physical and biological. Fortunately the problem is tractable in that the fundamental processes appear to be reasonably understood and can be described by relatively simple mathematical expressions into which values of parameters can be inserted. The present inadequacies arise more from a lack of data for some of these parameters,as follows.

Gas phase mass transfer coefficients and Henry's Law Constants for many reasonably volatile and soluble compounds between air and distilled water are adequately quantified for most environmental calculation purposes. Liquid phase mass transfer coefficients can be predicted within a factor of two for most conditions, an accuracy acceptable for many purposes such as determining whether or not volatili-zation rates are significant in comparison to other processes. The effects of surface films are poorly understood but in most unpolluted waters the effect is probably small. In heavily polluted waters, it is probably not necessary to achieve a high degree of accuracy. The major uncertainties are the values of H for many low solubility, low vapour pressure compounds and the effects of suspended and dissolved materials, particularly adsorbants.

There is a need to develop simple, fast, in situ methods of determining volatiliza-tion rates of pollutants of interest for various water bodies. Limnocorrals, as used by Emerson (10) appear to be the most promising approach, although the effect of their interference with horizontal flow and diffusion must be better quantified. For analytical purposes the use of radio-labelled compounds seems the logical approach. The appeal of such a method lies in its ability to measure volatiliza-rates directly and thus confirm the validity of calculations based on predicted values of kp, kj and H. In addition, it could help to elucidate the time dependence effect (diurnal, day-to-day and seasonal) of volatilization rates and thus

184 D. Mackay

elucidate how this effect should be taken into account in modelling or calculating average or annual fluxes.

ACKNOWLEDGEMENTS

The author is indebted to the National Research Council of Canada, Imperial Oil Limited and Environment Canada for financial support of work described in this paper.

REFERENCES

1. P.S. Liss, "Processes of gas exchange across an air-water interface", Deep Sea Res. 20, 221 (1973).

2. P.S. Liss, "The exchange of gases across lake surfaces", Proc. First Specialty Conf. on Atmospheric Contribution to the Chemistry of Lake Waters 88, Intern. Assoc. Great Lakes Res. (1975).

3. P.S. Liss and P.G. Slater, "Flux of gases across the air-sea interface", Nature 247, 181 (1974).

4. D. Mackay and P.J. Leinonen, "Rate of evaporation of low solubility contaminants from water bodies to atmosphere", Envir. Sei. Technol. 9, 1178 (1975).

5. R. Sutherland, "Determination of Henryfs Law Constants" M.A.Sc. Thesis, University of Toronto (1977).

6. E.R. Holley, "Turbulence measurements near the free surface of an open channel flow", Water Resources Res. 6, No. 3 960 (1970).

7. A. Pochmursky, "The estimation of vertical thermal eddy diffusivity from temperature profiles in water", B.A.Sc. Thesis, University of Toronto (1977).

8. W.B. Langbein and W.H. Durum, "The aeration capacity of streams", U.S. Geological Survey Circular 542, Washington D.C. (1967).

9. G.T. Csanady, "Turbulent Diffusion in the Environment", Reidel, Holland(1973)·

10. S. Emerson, "Gas exchange in small Candian Shield lakes", Limnol. and Oceanog. 20, 754 (1975).

11. W.S. Broecker and T.H. Peng, "The vertical distribution of radon in the Bomex area", Earth and Plant Sei. Lett. 11, 99 (1971).

12. S. Emerson, "Chemically enhanced CO« gas exchange in an entrophic lake: A general model", Limnol. and Oceanog., 20, 743 (1975).

13. W.S. Broecker and T.H. Peng, "Gas exchange between air and sea", Tellus 24, 21 (1974).

14. D.L. Thurber and W.S. Broecker, "The behaviour of radiocarbon in the surface waters of the Great Basin", Nobel Sympos. 12, 379, Wiley New York

15. Y. Cohen, "Exchange-processes across a free air-water interface", M.A.SC. Thesis, University of Toronto (1976).


REFERENCES

16. R.E. Treybal, "Mass transfer operations", 2nd Edn., McGraw Hill, N.Y. (1968).

17. E.C. Tsivoglou, "Tracer Measurement of stream reaeration", Fed. Water Poll. Contr. Admin., Div. of Tech. Services, U.S. Dept. of Interior, Washington, D.C. (1967).

18. D.M. Whelpdale, and R.W. Shaw, "Sulphur dioxide removal by turbulent transfer over grass, snow and water surfaces*1, Tellus, 26, 196 (1974).

19. P.S. Liss, "Chemistry of the sea surface microlayer", Chemical Oceanog., 2 J.P. Riley and G. Skirrow Eds. 193, Academic Press, London (1975).

Microbial Transformations of Aromatic Pollutants

DAVID T. GIBSON

Department of Microbiology, The University of Texas at Austin, Austin, Texas 78712

ABSTRACT

Over geological time microorganisms have evolved catabolic enzyme systems for the metabolism of naturally-occurring aromatic compounds and laboratory studies have revealed certain general features of aromatic metabolism. Thus, oxygen plays a central role in the hydroxylation and fission of the aromatic nucleus. Bacteria incorporate both atoms of molecular oxygen into aromatic hydrocarbons to form ois-dihydrodiols which then undergo further oxidation to catechols. In contrast mam-mals and fungi incorporate one atom of oxygen into aromatic hydrocarbons to form reactive arene oxides which undergo enzymatic hydration to form dihydrodiols with a trans-stereochemistry. The presence of two hydroxyl groups on the aromatic nucleus is a prerequisite for enzymatic fission of the ring. The hydroxyl sub-stituents may be ortho or para to each other. Subsequent metabolic sequences vary depending on the organism and the site of ring cleavage. Molecules synthe-sized by man are degraded if they bear a structural relationship to naturally-occurring compounds. These features are discussed in relation to the microbial transformations of PCBfs and chlorinated benzoic acids.

INTRODUCTION

Webster's dictionary defines pollutants as those components of manmade waste that contaminate the environment (1). This definition adequately describes those as-pects of pollution that are visible to the eye. Thus, the eutrophication of lakes and rivers is due, in many instances, to the introduction of large quantities of organic or inorganic nutrients. Another example is the foam that used to persist on rivers after the introduction of branched chain alkylbenzene sulfonates. These compounds were originally designed as detergents. In contrast, there are many com-pounds that persist in aquatic environments that give little indication as to their presence. Such molecules rarely cause immediate pertubations of the environment although concern for long term effects is a subject that must be investigated. Examples of this type of pollution include such molecules as DDT and the poly-chlorinated biphenyls (PCBTs). The rapid advances that have been made in analyti-cal techniques have made possible the detection of minute amounts of environmental pollutants. Most of these compounds are synthesized by man and are of relatively recent origin.

The benzenoid nucleus has proved particularly useful as the basis for the synthe-sis of a large number of chemicals that ultimately find their way into the environ-

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J88 D. T. Gibson

ment. Many insecticides, herbicides, drugs, explosives, solvents, detergents, etc. are aromatic in terms of their molecular structure. In addition, certain aromatic hydrocarbons are considered as environmental pollutants. There are indications that certain living organisms can synthesize polycyclic aromatic hydrocarbons (2). However, it appears that most of these compounds are produced by the thermal al-teration of organic matter. Sources of these compounds in the environment include high temperature pyrolysis due principally to the activities of man, medium temp-erature pyrolysis due to fires; these may be caused by man or occur naturally, and sedimentary diagenesis which leads to the production of fossil fuels (3). The num-bers and types of polycyclic hydrocarbons in the environment remain largely unknown (4). Nevertheless, interest in these molecules is due to the fact that many of them are mutagens and some can cause cancer in experimental animals (5). In fact Blumer has speculated that the environmental distribution of polycyclic hydrocar-bons may have played a role in the evolution of living organisms (6).

The microbial transformation of aromatic compounds has received wide attention over the past two decades. Microorganisms are the principal agents responsible for the recycling of carbon in nature. Thus, it is not surprising to find a variety of organisms that will degrade many different compounds. Hegeman has pointed out that chemical and biological evolution have taken place over a period of four billion years (7). During this time microorganisms have had the opportunity to develop the requisite enzyme systems for the degradation of naturally-occurring aromatic molecules. The latter include aromatic acids, phenols, lignin components and cer-tain halogenated compounds. On the other hand, the last two decades have seen the chemical synthesis of numerous aromatic structures that ultimately find their way into the environment. The capacity of the microbial flora to degrade such mole-cules will be dependent on the relationship of their chemical structure to naturally-occurring compounds (8).

Studies on the fate of aromatic pollutants in the environment are extremely complex. In addition to microbial transformation other environmental factors have to be con-sidered. These include, volatilization, photodecomposition, adsorption, solubili-zation and the complex interactions between many different organisms and organic compounds. Most analytical techniques will only permit studies on the disappear-ance or movement of the parent compound. In contrast, laboratory studies with pure strains of microorganisms and individual compounds gives information at the molecu-lar level but suffers from the criticism that such reactions cannot be guaranteed to occur in the environment. Nevertheless, as Dagley points out, the study of the degradation of pollutants is linked to both the mechanisms of induction and the modes of action of those enzymes employed for degrading natural products. Conse-quently, in laboratory studies on the degradation of pollutants there will come a time when the experimental findings must be interpreted within the framework of knowledge of microbial metabolism in general (8). A logical extension of this philosophy is that the same reasoning should also be applied, where possible, to the fate of pollutants in the environment. The following discussion attempts to show how laboratory studies can be used to determine the biodegradability of cer-tain classes of aromatic compounds. No attempt has been made to cover all aspects of this subject since several excellent reviews are available (9,10,11,12). Par-ticular emphasis is placed on studies in the author's laboratory on the microbial transformations of aromatic hydrocarbons and related compounds.

GENERAL FEATURES OF AROMATIC METABOLISM

For many years it was assumed that the degradation of aromatic compounds was de-pendent on the presence of atmospheric oxygen. The importance of enzymatic oxygen fixation into the benzenoid nucleus was first demonstrated by Mason (13) and Hayaishi (14). Mason showed that a phenolase complex would convert 3,4-dimethyl-phenol to 4,5-dimethylcatechol and proved that the oxygen in the incoming hydroxyl

Microbial Transformations of Aromatic Pollutants 189

group was derived exclusively from molecular oxygen. Hayaishi demonstrated that the enzyme pyrocatechase would oxidize catechol to eis-eis-muconlc acid; a reac-tion in which both atoms of molecular oxygen were incorporated into the ring fis-sion product. These reactions which are shown in Fig. 1 not only spawned numerous

- ^ 18 l/20g *CH3

CH3

4,5-Dimethylcatechol

a r0 H , 8o2 ^ c , 8 (

II *- L· ^ ιβ ,

eis, eis-Muconic acid Fig. 1. Examples of enzymatic oxygen fixation.

studies on the mechanisms of enzymatic oxygen fixation they proved that hydroxyla-tion and ring fission are important oxygenase reactions in the dissimilation of a variety of aromatic compounds. Such investigations were really initiated in 1932 when Happold and Key described the isolation of a Vibrio species that would degrade phenol, and meta- and para-hydroxybenzoate (15). Subsequent studies by Evans and his associates in Wales, Kilby in England and Stanier and his colleagues in the USA paved the way for the discovery of enzymatic oxygen fixation described above and and also for the elucidation of the 3-ketoadipate pathway which is used by certain strains of bacteria for the degradation of catechol and protocatechuic acid (12). Several pathways for the aerobic metabolism of dihydroxy aromatic compounds are now known (see below). In almost all cases the investigations were performed with organisms that could utilize the aromatic compound as a source of carbon and energy for growth.

In 1960 Leadbetter and Foster introduced the technique of co-oxidation which was based on the observations that many microorganisms have the ability to oxidize a variety of hydrocarbons even though they cannot use them as growth substrates (16). Thus, co-oxidation refers to the oxidation of nongrowth hydrocarbons when they are present as cosubstrates in a medium in which one or more different hydrocarbons are furnished for growth. This technique widened considerably the numbers of aro-matic compounds that are known to be metabolized by microorganisms and was elegantly developed by Raymond and his associates (17). Some examples of co-oxidation are shown in Fig. 2.

Horvath and Alexander (18) noted that "numerous microorganisms are able to metabo-

190 D. T. Gibson

GROWTH SUBSTRATE COSUBSTRATE

CH^CH^

CH3(CH2 )|4 CH3

PRODUCT

CH2COOH

CH3(CH2 )|4CH3

CH3

X^COOH L^COOH

CH

Fig. 2. Examples of co-oxidation.

lize substrates that they are unable to use either as a source of energy or of one of the elements necessary for proliferation." They termed this phenomenon cometabo-lism and showed that resting cell suspensions of a benzoate-grown Arthrobacter sp. would quantitatively oxidize tf?-chlorobenzoate to 4-chlorocatechol. The bacterium would not grow with tf?-chlorobenzoate. Apparently the enzymes responsible for the conversion of benzoate to catechol can tolerate a halogen substituent in the meta-position whereas the catechol oxygenase in this organism cannot metabolize 4-chloro-catechol. Similar results were reported earlier by Dagley and Patel (19) for the partial metabolism of p-cresol analogues by cells of Pseudomonas species that were grown with p-cresol as the sole source of carbon. Some of the products accumulated by this organism are given in Fig. 3. They attributed their results to the broad specificity of some of the enzymes responsible for the metabolism of p-cresol.

Studies such as those described above have revealed certain general features that pertain to the aerobic degradation of aromatic compounds. Dihydroxylation of the aromatic nucleus is a prerequisite for fission of the ring. The hydroxyl groups may be ortho to each other as in catechol or para to each other as in gentisic acid. As a result the biodegradation of aromatic compounds may be considered from two aspects; preparation for ring fission and the transformations that occur after the cleavage of the aromatic nucleus. The persistence of aromatic pollutants or their transformation products in the environment is probably a reflection of their inability to undergo these types of reactions.

PREPARATION FOR RING FISSION

The mechanisms used for the introduction of hydroxyl groups into benzenoid com-


pounds depend on the substituents that are already present on the molecule and also on the organism effecting the transformation. For many years it was assumed that hydroxylation reactions were catalyzed by monooxygenases according to the following reaction sequence:

RH + XH2 + 02 -* ROH + X + H20

RH = Substrate X = Electron Donor

This type of reaction may be illustrated by the conversion of phenol (20) and p-hydroxybenzoate (21) to catechol and protocatechuic acid respectively. Thus one atom of molecular oxygen is incorporated into the substrate and the other atom is converted to water.

9H3 COOH

Φ OH

CH,

OH

RING FISSION

CH, COOH

<r —- <r OH OH

CH. COOH

σ —-—-—-tf OH OH

Fig. 3. Products accumulated from p-cresol analogues by cells of a Pseudomonas sp.

J92 D. T. Gibson

A different mechanism has been shown to operate for the formation of catechol and substituted catechols from molecules that do not already contain a hydroxy sub-tituent. Benzene, toluene and ethylbenzene will each serve as a growth substrate for a strain of Pseudomonas putida. A mutant strain of this organism was isolated that would oxidize these aromatic hydrocarbons to cts-dihydrodiols (22,23,24). When the formation of c£s-l,2-dihydroxy-l,2-dihydrobenzene (cis-benzenedihydrodiol) was carried out in the presence of isotopic oxygen both atoms of molecular oxygen were incorporated into the molecule (22). The parent organism was shown to contain a dehydrogenase that oxidized the <?£s-dihydrodiols to catechols. The reactions catalyzed by this organism for benzene metabolism are shown in Fig. 4. It is of

Benzene

k^4r° Hypothetical

Dioxetane

a40H _ £ 0 H

H

<r/s-l,2-Dihydroxy-l,2-

dihydrobenzene

Further Degradation

Fig. 4. Initial reactions in the degradation of benzene by Pseudomonas putida.

interest to note that alkyl substituents up to butyl can be metabolized to catechols which can undergo ring fission (25). In contrast, compounds such as chloro, bromo, and iodobenzene as well as p-xylene and p-chlorotoluene are converted through ois-dihydrodiols to their respective catechols which are resistant to further degrada-tion (26,27). Also we have noticed that 3-chlorocatechol will cause the accumula-tion of catechols from benzene and toluene. Under laboratory conditions catechols that are catabolic intermediates rarely accumulate in the culture medium. At this time the mechanism of inhibition by 3-chlorocatechol has not been elucidated. The metabolism of monocyclic aromatic compounds through cis-dihydrodiols is now known to be a common reaction in the bacterial degradation of aromatic compounds. Table 1 lists those reactions where a ais relative stereochemistry has been inferred or established.

Bacteria also metabolize polycyclic aromatic hydrocarbons through cis-dihydrodiol intermediates. Naphthalene and anthracene are oxidized to ois-l(R),2(S)-dihydroxy-1,2-dihydronaphthalene (35) and ois-l(R),2(S)-dihydroxy-l,2-dihydroanthracene (36) respectively. A Beijerinokia strain oxidizes phenanthrene through c£s-3,4-dihy-droxy-3,4-dihydrophenanthrene and a cis-dihydrodiol is also formed at the Im-position (37). At this time no organisms have been isolated that will grow with polycyclic hydrocarbons that contain more than three aromatic rings. However, the Beijerinokia strain mentioned above will grow with biphenyl and phenanthrene. When cells of this organism are induced with these substrates and then incubated with either benzo(a)pyrene or benzo(a)anthracene the latter compounds are converted to acid products. The structural relationship between phenanthrene biphenyl and larger polycyclic hydrocarbons are shown in Fig. 5. Beijerinokia B8/36 is a mutant organism that oxidizes biphenyl and phenanthrene to cis-dihydrodiols at the 2,3- and 3,4-position respectively (28,37). If the enzymes induced by these sub-strates also attack larger polycyclic aromatic compounds one would predict the formation of ois-9,10-dihydroxy-9,10-dihydrobenzo(a)pyrene and c*£s-l,2-dihydroxy-1,2-dihydrobenzo(a)anthracene. These compounds are indeed formed (38). However,


TABLE 1 Monocyclic aromatic hydrocarbons and related compounds that are oxidized to cis-dihydrodiols by bacteria

Substrate

Benzene Toluene Chlorobenzene p-Chlorotoluene p-Bromotoluene p-Fluorotoluene Ethylbenzene p-Xylene Biphenyl Benzoid Acid 3-Chlorobenzoic Acid

3,5-Dichlorobenzoic Acid 4-Chlorobenzoic acid 3,4-Dichlorobenzoic Acid

1 3-Methylbenzoic Acid 3,5-Dimethylbenzoic Acid

i 4-Methylbenzoic Acid 3-Fluorobenzoate 4-Fluorobenzoate 4-Trifluoromethylbenzoate 5-Amino-4-chloro-2-phenyl-3

(2H)pyridazinone a-Methylstyrene

Organism

Unidentified. Strain E Unidentified. S107B1

Ref.

(22) (23) (26) (26) (26) (26) (24) (27) (28) (29) (30) (29) (31) (31) (31) (31) (31) (31) (29) (29) (32)

(33) (34)

the technique of high pressure liquid chromatography led to the detection and identification of small amounts of other dihydrodiols. Benzo(a)pyrene also gave a dihydrodiol at the 7,8-position and benzo(a) anthracene was converted to dihydrodiols at the 8,9- and 10-11 positions (39). The further metabolism of these products has not been studied in detail. Some of the reactions utilized by bacteria for the conversion of aromatic compounds to catechols are shown in Fig. 6.

The initial reactions utilized by bacteria for the oxidation of aromatic hydro-carbons are different to those observed in higher organsims. Mammals oxidize these compounds to arene oxides which can then isomerize to form phenols, undergo conjugation with glutathione or react with water to form trans-dihydrodiols (40). Recently we have shown that fungi oxidize naphthalene in a manner analogous to that reported for liver microsomes (41). When Cunninghamella elegans was grown in the presence of naphthalene α-naphthol was shown to be the major metabolite. Other degradation products were 3-naphthol, trans-1,2-dihydroxy-l,2-dihydronaphthalene, 4-hydroxy-l-tetralone, 1,2- and 1,4-naphthoquinone (42). A proposed reaction sequence for the degradation of naphthalene by this organism is shown in Fig. 7.

RING FISSION

Catechol and protocatechuate are central intermediates in the degradation of a

Pseudomonas putida Pseudomonas putida Pseudomonas putida Pseudomonas putida Pseudomonas putida Pseudomonas putida Pseudomonas putida Pseudomonas putida Beijerinokia species Aloaligenes eutrophus Pseudomonas sp. Aloaligenes eutrophus Aloaligenes eutrophus Aloaligenes eutrophus Aloaligenes eutrophus Aloaligenes eutrophus Aloaligenes eutrophus Aloaligenes eutrophus Aloaligenes eutrophus Aloaligenes eutrophus Pseudomonas putida

194 D. T. Gibson

6 I

5'

2

»3

BIPHENYL

7 6 5

BENZO W3PYRENE

8 9 PHENANTHRENE

11 12

8 7 6

BENZO Bfl ANTHRACENE Fig. 5. Structural relationships between polycyclic aromatic hydrocarbons.

H

χ Ό Η

J _ Λ Λ 1 . COOH

XOH-RING

.OH FISSION CATECHOL

2H+C02

H Fig. 6. Catechol formation by bacteria.


nophtholent

< 1,2-nophtho-

I t n t oxidt

H * 0

<§& H

trans-1,2- dihydroxy -1 ,2- dihydronophtholtnt

1,2 - dihydroxy nophthol tn t

1,2-nophthoquinont

2-nophthol

OH

l-nophthol

\ OH

OH .

s 1,4-dihydroxy-

nophthaltnt

0 1,4-naphthoquinone

\ Ψ

0

βφ HÖH

4-hydroxy-l- tetrolone

Fig. 7. Proposed reaction sequence for the metabolism of naphthalene by Cunninghamella elegans.

variety of aromatic compounds (Table 2). Consequently the metabolism of these hy-droxylated intermediates has received considerable attention. In addition, 1,4-dihydric phenols such as homogentisic acid and quinol are typical intermediates in the degradation of phenylalanine and p-nitrophenol respectively. The differ-ent pathways utilized by bacteria for the degradation of phenolic compounds are beyond the scope of this article. They have been discussed in detail by Chapman (10), Ornston and Stanier (12) and Dagley (9). Some examples of the different kinds of ring-fission reactions that are catalyzed by bacteria are shown in Fig. 8.

196 D. T. Gibson

TABLE 2 Catechol and Protocatechuic acid as central intermediates in the degradation of various aromatic substrates.

Compounds Metabolized Through 1

CATECHOL

Mandelic acid Toluene Benzene Benzoic acid Naphthalene Phenol Anthranilic acid Tryptophan Salicyclic acid Naphthalene

PROTOCATECHUIC ACID

p-Hydroxymandelic acid p-Hydroxybenzoic acid p-Toluic acid w-Hydroxybenozic acid

Shikimic acid Quinic acid Vanillic acid Phenanthrene Phthalic acid

w-Cresol p-Cresol w-Nitrobenzoic acid p-Nitrobenzoic acid

Catechol

HOOCs^s^OH

Protocatechuic acid

COOH ,0H jr.—--Ö

a- Hydroxymuconic semialdthyd·

CCOOH COOH

Cftf,£/f-muconate

HOOC ^ . _ Λ Λ 1 1 %^S>COOH l ^ C O O H

ß-Carboxy- f/fr.g/5- muconate

ΗΟθγ^ΟΗ

o S l COOH ß-Hydroxy-γ- corboxy-

muconic semialdehyde COOH 1=0

COOH

. Formate or C02

acetaldehyde and pyruvate

Ketoadipate. Acetyl-CoA and Succinate

Formate or C 0 2

and 2 pyruvate

Fumarate and pyruvate

Gentisic acid Maleylpyruvic acid

Fig. 8. Ring-fission products formed from dihydroxylated aromatic compounds.


DEGRADATION OF AROMATIC POLLUTANTS

The above brief introduction to the metabolism of aromatic compounds by microorgan-isms provides valuable information pertaining to an understanding of the trans-formation of aromatic pollutants. In this context pollutants may be considered as molecules that have been introduced into the environment in recent years by the activities of man. For examples, organochlorine compounds are used in vast quan-tities as solvents, lubricants, insulators, insecticides, herbicides, plasticizers etc. Some of these compounds are readily degraded by microorganisms while others appear to be resistant to microbial attack. If the tenets of microbial evolution mentioned previously are true, synthetic organochlorine compounds will be meta-bolized only if they will serve as substrates for pre-existing enzyme systems. In the absence of such enzymes the compounds will persist until microorganisms evolve the metabolic capability to deal with the new chemical structure (43).

Pre-existing enzyme systems fall into two categories. Those that have evolved for the degradation of halogen-free compounds which will fortuitously metabolize halo-genated analogues and those that have evolved specifically for the degradation of naturally-occurring halogenated compounds. Examples of both types of degradation may be seen in the metabolism of chlorinated biphenyls, and chlorinated benzoic acids.

DEGRADATION OF CHLORINATED BIPHENYLS

Since chlorinated biphenyls are relatively recent newcomers to the environment it is unlikely that microorganisms will have evolved the metabolic capability to degrade these compounds. If degradation is to occur at all it must be effected by pre-existing enzyme systems. It is easy to isolate microorganisms that will de-grade the unsubstituted parent molecule (44,45,46,47,48,28). Although the re-action sequence for the degradation of biphenyl has not been completely elucidated the available evidence supports the pathway shown in Fig. 9. Organisms that can utilize this pathway can also metabolize certain chlorinated biphenyl derivatives. However, there is no evidence that any of the enzymes involved in biphenyl degrad-ation can dehalogenate the chlorinated compounds. Consequently different organ-isms accumulate different halogenated intermediates. When chlorine substitution occurs in only one of the two aromatic rings, microorganisms preferentially metab-olize the unsubstituted aromatic nucleus. As a result chlorinated benzoic acids are produced (46,47,48). There are some indications that biphenyl-degrading or-ganisms will also metabolize 3-chloro- and 3,3f-dichlorobiphenyl past the chlorin-ated benzoic acid state (48). However, there are no indications that chloride is ever released from the degradation products. When chlorination occurs in both aromatic rings only those molecules with unsubstituted 2,3-positions appear to be degraded. Also chlorination at the 4,4f-positions leads to the accumulation of ring-fission products rather than chlorinated benzoic acids (48).

Some of the types of chlorinated metabolites that accumulate when different strains of microorganisms oxidize chlorinated biphenyl isomers are shown in Fig. 10. Although relatively little information is available, the evidence seems to indicate that the bacterial degradation of some of the components in PCB mixtures occurs fortuitously. Bacteria seem to require an unsubstituted 2,3-position in order to initiate the degradation of chlorinated biphenyls and these organisms do not have the capacity to dehalogenate either the parent compounds of the chlorinated de-gradation products. At this time there is no reason to assume the existence of bacteria that will metabolize the highly chlorinated isomers of biphenyl. The studies described above would predict that the 2,6,2f,6f- and 3,5,3f,5T-tetra-chlorobiphenyls would be particularly resistant to bacterial degradation. The possibility exists that molds and yeast may metabolize PCBfs more readily than bacteria. These organisms contain non-specific enzyme systems for the metabolism

198 D. T. Gibson

of a variety of aromatic substrates and the products formed are similar to those produced by higher organisms (42,49,50,51). Thus fungi have been shown to oxidize 4-chlorobiphenyl and 4,4T-dichlorobiphenyl to 4-chloro-4f-hydroxybiphenyl and 4.41

dichloro-3-hydroxybiphenyl respectively (52,53). It is conceivable that the variety of hydroxylated products formed from chlorinated biphenyls by higher organ-isms (53) will also serve as substrates for microbial degradation. Nevertheless, the potential exists for the accumulation in the environment of highly chlorinated biphenyls and chlorinated biphenyl metabolites. Clearly much more work needs to be done in this area.

5' 6' 6 5

OH OH

u °" -COOH

RING FISSION

Fig. 9. Proposed pathway for the degradation of biphenyl by bacteria.


-9 Cl

- ?

Fig. 10. Products formed from some chlorinated biphenyls by bacteria.

DEGRADATION OF CHLORINATED BENZOIC ACIDS

Certain chlorinated benzoic acids are fortuitously metabolized by the enzymes of benzoate metabolism. This sequence results in the conversion of benzoate to catechol. Organisms, grown with benzoate, can oxidize 3-chlorobenzoic acid either to 3-chlorocatechol (54,55) or 4-chlorocatechol (56,57). The halogenated catechols accumulate because they cannot serve as substrates for the ring-fission enzymes in these organisms. In contrast Knackmuss and his associates used a chemostat to enrich for an organism that would degrade 3-chlorobenzoate. The chemostat originally contained benzoate as the sole source of carbon and after one week of continuous operation the benzoate was gradually replaced with 3-chlorobenzoate. The change to the halogenated benzoate took two weeks and after a further five weeks a strain of Pseudomonas ftuoresoens was isolated that would grow with 3-chlorobenzoate (30). When this organism was grown with benzoate it was shown to contain enzymes that would convert 3-chlorobenzoate through the analogous diol intermediates to both 3- and 4-chlorocatechol. The catechol-l,2-oxygenase in this organism showed very little activity with the halogenated catechols. In contrast, when P. ftuoresoens was grown with 3-chlorobenzoate a second catechol-l,2-oxygenase was induced that would rapidly oxidize both 3- and 4-chlorocatechol to chlorinated muconic acids. Special enzymes were also present for the laconization and isomer-ization reactions that occur later in the 3-ketoadipate pathway (30,31,58,59). It appears that the organism has evolved the ability to synthesize at least two new enzymes that are capable of metabolizing both 3- and 4-chlorocatechol through a

200 D. T. Gibson

halogenated version of the ß-ketoadipate pathway (Fig. 11). In this sequence

oc

COOH ^ S ^ C O O H

9 Cl

c OH

COOH COOH

I OH

COOH COOH

_ >COOH

Y^COOH k^xCOOH

Fig. 11. Bacterial degradation of benzoic acid and 3-chlorobenzoic acid.

only the degradation of 3-chlorocatechol is shown and the mechanism of chlo-ride release has not been established. Nevertheless this is one example of of evolution of an enzyme system, possibly by gene duplication and modifica-tion, to permit the biodegradation of a halogenated aromatic substrate.


Another example could be the degradation of chlorinated phenoxyacetic acid herbidices (60,61,62,63,64,65). Organisms could have evolved the enzymes necessary for the specific metabolism of these compounds. If this is the case the evolutionary changes would have occured over the past thirty years. Alternatively, microorganisms have been exposed to natural products such as 2,4-dichlorophenol for a considerably longer period of time (66). Consequent-ly, organisms that can degrade 2,4-dichlorophenol would only have to evolve the ability to remove the ether sidechain in order to be able to metabolize 2,4-dichlorophenoxyacetic acid.

ACKNOWLEDGEMENTS

Some of the studies described in this manuscript were supported in part by grants ES-00537, awarded by the National Institute of Environmental Health Sciences; 1 ROI CA19078 awarded by the National Cancer Institute, DHEW; F-440 from the Robert A. Welch Foundation; the Office of Naval Research, Micro-biology Program, Naval Biology Project, under contract N00014-76-C-0102, NR205-008 and Contract N01 CP 33384 awarded by the National Cancer Institute, DHEW. The author was a recipient of Career Development Award 1 K04 ES-70088 from the Institute of Environmental Health Sciences, DHEW.

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202 D. T. Gibson

T. Gibson, The microbial oxidation of aromatic hydrocarbons, Crit. Rev. Microbiol. 1, 199 (1971). Y. Stanier and L. N. Ornston, The 3-ketoadipate pathway, Adv. Microbial Physiol. 9, 89 (1973). S. Mason, W. L. Fowlks, and E. Peterson, Oxygen transfer and electron transport by the phenolase complex, J. Amer. Chem. Soc. 77, 2914 (1955). Hayaishi, M. Katagiri and S. Rothberg, Mechanism of the pyrocatechase reaction, J. Amer. Chem. Soc. 77, 5440 (1955). C. Happold and A. Key, The bacterial purification of gas works liquors. The action of the liquors on the bacterial flora of sewage, J. Hy g. (Britain) 32, 573 (1932).

R. Leadbetter and J. W. Foster, Bacterial oxidation of gaseous alkanes, Arch. Mikrobiol. 35, 92 (1960). L. Raymond and V. W. Jamison, Biochemical activities of Nocardia, Adv. Appl. Microbiol. 14, 93 (1971). S. Horvath and M. Alexander, Cometabolism: a technique for the accumulat-ion of biochemical products, Can. J. Microbiol. 16, 1131 (1970).

Dagley and M. D. Patel, Oxidation of p.-cresol and related compounds by a pseudomonasj Biochem J. 66, 227 (1957). C. Evans, Oxidation of phenol and benzoic acid by some soil bacteria, Biochem. J. 41, 373 (1947). P. Sleeper and R. Y. Stanier, The bacterial oxidation of aromatic compounds I. Adaptive patterns with respect to polyphenolic compounds, J. Bacteriol. 59, 117 (1950). T. Gibson, G. E. Cardini and R. E. Kallio, Incorporation of oxygen-18 into benzene by Pseudomonas putida* Biochemistry 9, 1631 (1970). T. Gibson, M. Hensley, H. Yoshioka and T. J. Mabry, Formation of (+)-cis-2,3-dihydroxy-l-methylcyclohexa-4,6-diene from toluene by Pseudomonas putida, Biochemistry 7, 1626 (1970). T. Gibson, B. Gschwendt, W. K. Yeh and V. M. Kobal, Initial reactions in the oxidation of ethylbenzene by Pseudomonas putida. Biochemistry 12, 1520 (1973). T. Gibson, J. R. Koch and R. E. Kallio, Oxidative degradation of aromatic hydrocarbons by microorganisms, I. Enzymatic formation of catechol from benzene, Biochemistry 7, 2653 (1968). Ziffer, K. Kabuto, D. T. Gibson, V. M. Kobal and D. M. Jerina. The absolute stereochemistry of several cis-dihydrodiols microbially produced from substituted benzenes, Tetrahedron (in press, 1977). T. Gibson, V. Mahadevan and J. F. Davey, Bacterial metabolism of para-and meta-xylene: oxidation of the aromatic ring, J. Bacteriol. 119, 930 (1974).

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(ID

(12)

(13)

(14)

(15)

(16)

(17)

(18)

(19)

(20)

(21)

(22)

(23)

(24)

(25)

(26)

(27)

(28)

(29)

(30)

(3 i ;

(32;

(33

D.

R.

H.

0 .

F.

E.

R.

R.

S.

W.

B.

D.

D.

D.

D.

H.

D.

D.

A.

E.

) H.

) J .

) E.


(34) T. Omori, Y. Jigami and Y. Minoda, Microbial oxidation of a-methylstyrene and B-methylstyrene, Agr. Biol. Chem. 38, 409 (1974).

(35) A. M. Jeffrey, H. J. C. Yeh, D. M. Jerina, T. R. Patel, J. F. Davey and D. T. Gibson, Initial reactions in the oxidation of naphthalene by Pseudomonas putida, Biochemistry 14, 575 (1975).

(36) M. N. Akhtar, D. R. Boyd, N. J. Thompson, M. Koreeda, D. T. Gibson, V. Mahadevan and D. M. Jerina, Absolute stereochemistry of the dihydro-anthracene-c£s-and-t:ratts-l,2-diols produced from anthracene by mammals and bacteria, J. Chem. Soc. 2506 (1975).

(37) D. M. Jerina, H. Selander, H. Yagi, M. C. Wells, J. F. Davey, V. Mahadevan and D. T. Gibson, Dihydrodiols from anthracene and phenanthrene, J. Am. Chem. Soc. 98, 5988 (1976).

(38) D. T. Gibson, V. Mahadevan, D. M. Jerina, H. Yagi and H. J. C. Yeh, Oxidation of the carcinogens benzo(a)pyrene and benzo(a)anthracene to dihydrodiols by a bacterium, Science 189, 295 (1975).

(39) D. T. Gibson and D. M. Jerina (unpublished observations). (40) D. M. Jerina and J. W. Daly, Arene oxides: a new aspect of drug metabolism,

Science 185, 573 (1974). (41) D. M. Jerina and J. W. Daly, B. Witkop, P. Zaltzman-Nirenberg and S.

Udenfriend, 1,2-Naphthalene oxide as an intermediate in the microsomal hydroxylation of naphthalene, Biochemistry, 9, 147 (1969).

(42) C. E. Cerniglia and D. T. Gibson, Metabolism of naphthalene by Cunninghamella eleganSy Appl. and Environ. Microbiol. (1977, in press).

(43) P. J. Chapman, Microbial degradation of halogenated compounds, Biochem. Soc. Transactions, 4, 16 (1976).

(44) D. Lunt and W. C. Evans, The microbial metabolism of biphenyl, Biochem. J. 118, 54P (1970).

(45) D. Catelani, C. Sorlini and V. Treccani, The metabolism of biphenyl by Pseudomonas putida, Experientia, 27, 1173 (1971).

(46) M. Ahmed and D. D. Focht, Degradation of polychlorinated biphenyls by two species of Achromobacter, Can. J. Microbiol. 19, 47 (1973).

(47) T. Ohmori, T. Ikai, Y. Minoda and K. Yamada, Utilization of polyphenyl and Polyphenyl-related compounds by microorganisms, Agr. Biol. Chem. 37, 1559 (1973).

(48) K. Furukawa and F. Matsumura, Microbial metabolism of polychlorinated biphenyls. Studies on the relative degradability of polychlorinated biphenyl compon-ents by Alkaligenes sp. Agr. Food Chem. 24, 251 (1976).

(49) J. P. Ferris, M. J. Fasco, F. L. Stylianopoulou, D. M. Jerina, J. W. Daly and A. M. Jeffrey, Monooxygenase activity in Cunninghamella bainievi: Evidence for a fungal system similar to liver microsomes, Arch. Biochem. Biophys. 156, 97 (1973).

(50) R. V. Smith and J. P. Rosazza, Microbial models of mammalian metabolism. Aromatic Hydroxylation, Arch. Biochem. Biophys. 161, 551 (1974).

(51) B. J. Auret, D. R. Boyd, P. M. Robinson, C. G. Watson, J. W. Daly and D. M. Jerina, The NIH shift during the hydroxylation of aromatic substrates by fungi, Chem. Commun. 1585 (1971).

(52) P. R. Wallnöfer, G. Engelhardt, S. Safe and 0. Hutzinger, Microbial hydroxy-lation of 4-chlorobiphenyl and 4,4' dichlorobiphenyl, Chemosphere, 2, 69 (1973).

(53) G. Sundström, 0. Hutzinger and S. Safe, The metabolism of chlorobiphenyls-a review, Chemosphere, 5, 267 (1976).

(54) A. Ichihara, K. Adachi, K. Hosokawa and Y. Takeda, The enzymatic hydroxy-lation of aromatic carboxylic acids; substrate specificities of anthra-ilate and benzoate oxidases, J. Biol. Chem. 237, 2296 (1962).

(55) N. Walker and D. Harris, Metabolism of 3-chlorobenzoic acid by Azotobaotev species, Soil Biol. Biochem. 2, 27 (1970).

(56) R. S. Horvath and M. Alexander, Co-metabolism of w-chlorobenzoate by an Arthrobaoter, Appl. Microbiol. 20, 254 (1970).

204 D. T. Gibson

(57) J. R. Spokes and N. Walker, Chlorophenol and chlorobenzoic acis co-metab-olism by different genera of soil bacteria, Arch. Microbiol. 96, 125 (1974).

(58) H. J. Knackmuss, M. Hellwig, H. Lackner and W. Otting, Cometabolism of 3-methylbenzoate and methylcatechols by a 3-chlorobenzoate utilizing Pseudomonas: accumulation of (+)-2,5-dihydro-4-methyl- and (+)-2,5-dihydro-2-methyl-5-oxo-furan-2-acetic acid, Eur. J. Appl. Microbiol. 2, 267 (1976).

(59) H. J. Knackmuss, W. Beckman, E. Dorn and W. Reineke, On the mechanism of the biological persistance of halogenated and sulfonated aromatic hydrocarbons, Zbl. Bakt. Hyg. 162, 127 (1976).

(60) -J. M. Tiedje, J. M. Duxbury, M. Alexander and J. E. Dawson, 2,4-D Metabolism: pathway of degradation of chlorocatechols by Arthrobaater sp.. J. Agric. Food Chem. 17, 1021 (1969).

(61) J. M. Tiedje and M. Alexander, Enzymatic cleavage of the ether bond of 2,4-dichlorophenoxyacetate, J. Agric. Food Chem. 17, 1080 (1969).

(62) J. M. Bollag, G. G. Briggs, J. E. Dawson and M. Alexander, 2,4-D Metabolism, enzymatic degradation of chlorocatechols, J. Agric. Food Chem. 16, 829 (1968).

(63) W. C. Evans, B. S. W. Smith, H. N. Fernley and J. I. Davies, Bacterial metabolism of 4-chlorophenoxyacetate, Biochem. J. 122, 509 (1971).

(64) J. K. Gaunt and W. C. Evans, Metabolism of 4-chloro-2-methylphenoxyacetate by a soil pseudomonad. Preliminary evidence for the metabolic pathway, Biochem. J. 122, 519 (1971).

(65) J. K. Gaunt and W. C. Evans, Metabolism of 4-chloro-2-methylphenoxyacetate by a soil pseudomonad. Ring-fission, lactonizing and delactonizing enzymes, Biochem. J. 122, 533 (1971).

(66) K. Ando, A. Kato and S. Suzuki, Isolation of 2,4-dichlorophenol from a soil fungus and its biological significance, Biochem. Biophys. Res. Commun. 39, 1104 (1970).

Photochemical Trans formation of Pollutants in Water

GÖRAN SUNDSTRÖM* and L. OCTAVIO RUZO**

*Special Analytical Laboratory, National Swedish Environment Protection Board, Wallenberg Laboratory, S-106 91 Stockholm, Sweden

**Division of Entomology and Parasitology, University of California, Berkeley, California 94720, USA

The photochemistry of pesticides and other xenobiotic compounds has rapidly become an integrated part in studies of the environmental de-gradation of pollutants nowadays present in rivers, lakes and oceans all over the world (1). At present some regulatory agencies require information as to the photostability and photoproduct identity of commercial chemicals prior to registration while the photochemical properties of substances already in use are being studied with renew-ed interest. We are thus concerned with the photodegradation of both actual and potential pollutants. It is also important to identify photoproducts formed and to investigate their toxicity since photo-transformations do not necessarily render a compound harmless. For example the photoproducts of aldrin, dieldrin and heptachlor have all been reported to be more toxic to several animal species than the pa-rent compounds (2). For producers it should also be of importance to assess the photostability of new compounds which makes kinetic stu-dies a necessary part of any photochemical investigation.

Initially most photodegradation studies have to be performed in labo-ratory model systems. Such studies have often been criticized because these systems do not reflect actual environmental conditions ( sol-vents, wavelengths used etc.) and results from such experiments have sometimes been erroneously extrapolated to natural conditions. How-ever, the results obtained with these systems, such as photoproduct identity and degradation rates are necessary as guidelines for envi-ronmental testing since severe analytical problems otherwise would arise. The recent identification of photodegradation products of some pesticides in environmental material have also confirmed that photo-decomposition of pollutants does occur under natural conditions, e.g. mirex (3), DDE (4), 2,4,5-T (5), dieldrin (6).

The present discussion will especially deal with recent reports on the phototransformations of pollutants from which extrapolations to actual aqueous environmental conditions, i.e. aqueous media, wave-lengths above 290 rim etc., is at least partially valid. Of course, an exact definition of an aqueous system with regard to pollutant photo-chemistry is difficult to formulate since the majority of the substan-ces of concern are organic compounds and often exhibit low solubili-ties in water. They therefore tend to accumulate in hydrophobic envi-ronments such as oil layers and cuticular waxes (7) or become adsorbed

205

206 G. Sundstrom and L. Octavio Ruzo

on suspended particulate matter and sediments. Photochemical reac-tions in the presence of water are also possible in the vapour phase where high humidity may cause substances to come in contact with wa-ter, such as may be the case in the marine aerosol. As will be shown in some examples below both product distribution and reaction rates will be highly dependent on both the micro-environment and the phy-sical state of the pollutant under study.

The determining process for phototransformations to occur is the ab-sorption by a molecule of a photon of the required energy. The absor-bing molecule is thereby excited to an electronic state with excess energy which can be converted to other forms of energy such as lumi-nescence (phosphorescence, fluorescence) or chemical reactions. When the sunlight reaches the surface of the earth its high energy compo-nents above ca. 290 nm have been removed by absorption in the ozone layer of the athmosphere. Still, the energy available in the residu-al light - ca. 95 Kcal/mol for 300 nm - is sufficient to break most covalent bonds, such as C-C (80-90 Kcal), C-H (95-100 Kcal) and C-Cl bonds (80 Kcal). It has been noted, however, that the quantum flux density below 290 nm can be as high as 1016 photons cm"1 month"**1 so over extended periods of time the lower wavelengths can be effective in causing photoreactions as well (8). It has thus been shown that a number of pesticides which, on the basis of their absorption spectra (9), absorb no or very little solar energy still are photolysed by sunlight (i.e. cyclodiene insecticides, urea herbicides and pyrethro-ids, see also above). As discussed below other reasons for the photo-reactions of such compounds are also possible.

The effect of latitude and season on the photolysis half-life, as exemplified with 2,4-D butoxyethyl ester, has been computed by Zepp et al. (10). At the northern latitude of 50° the midday half-life of this compound spans from about 250 h (july-august) to about 4 000 h (february-march), Fig. 1.

midday halflife, h

4000

3000H

2000

1000

time of year

Fig. 1. Dependence of 2,4-D butoxyethyl ester pho-tolysis half-life upon season and northern latitu-

de (from Ref. 10).

Photochemical Transformation of Pollutants in Water 207

Another manner in which photochemical reactions of compounds that do not exhibit absorption bands above 300 nm may occur, is by transfer of energy (sensitisation), from a molecule in its proximity that has absorbed radiation. The first molecule then acts as a quencher and itself becomes excited. Most known cases of sensitisation occur from the long-lived excited states, normally the triplet. It should be kept in mind that triplets are quenched by oxygen which means that sensitisation might be less efficient under some natural conditions. Several naturally occurring compounds such as chlorophylls and flavo-noid derivatives have been shown to be able to sensitise the photo-reactions of pesticides (11). The simultaneous application of a pes-ticide and a suitable sensitiser may then be used to accellerate the photoalterations or photodegradation of the pesticide such as is the case when rotenone and its photoproduct rotenolone are sprayed to-gether with some chlorinated cyclodiene pesticides (11) or triphenyl-amine with DDT (11,12).

An interesting case of photosensitisation from a toxicological point of view is the riboflavin-5"-sodium phosphate (FMN) sensitised for-mation of the chloracnegenic product 3,3",4,4^-tetrachloroazobenzene (13) from 3,4-dichloroaniline upon photolysis in water (14). The lat-ter compound is a common degradation product of many herbicides of the urea, amide and carbamate groups (Fig. 2).

Cly*ryNH2 FMN . C l V ^ V N = N > ^ V C l

Fig. 2. Photochemical formation of azo-derivatives from 3,4-dichloroaniline (from Ref. 14).

In addition to sensitising phenomena batochromic shifts in the absorp-tion spectrum of a compound upon adsorption on a surface as compared to the dissolved compound may be responsible for a change in photo-chemical reactivity. For example, it has been reported that the ab-sorption maximum of paraquat, when adsorbed on a surface, is increa-sed to 275 nm (15). Similar behaviour was reported for photodieldrin (193 nm in hexane, 264 nm adsorbed on silica gel) (16). The product distribution pattern is also sometimes changed when the compounds are adsorbed onto surfaces. An example is the preferential phototransfor-mation of aldrin to dieldrin, instead of photo-aldrin, when adsorbed on silica (17).

There is now a rapidly growing knowledge on the photochemical behavi-our of various types of pesticides and especially of the chlorinated pollutants. In the case of halogonated hydrocarbons the field is do-minated by reports of free radical intermediates and electron trans-fer processes in the photochemical reactions (18). Photolysis in the aqueous phase have not been studied as thoroughly as that in organic solvents, mainly due to solubility problems as mentioned above. There-


fore, certain aspects that are part of every photochemist^s backgro-und have to be re-evaluated before extrapolating results to water so-lutions. Oxygen, for example, which has a low triplet energy and thus acts as a very efficient triplet quencher, may play a lesser role in water since its solubility in water is lower than in most organic solvents.

Inorganic ions can also act as quenchers in aqueous media (19,20) by energy transfer mechanisms while their role in organic media may be negligible. Halogen atom salts such as lithium bromide and lithium iodide have recently been found to decrease the photoreaction rates of some halogenated aromatic compounds by increasing the decay rate from the triplet to the ground state by perturbation of the singlet--triplet intersystem crossing probability (21), Table 1.

TABLE 1 Effect of degradation rates

: lithium of some !

bromide and halogenated

iodide on the photo-aromatic compounds

(from Ref. 21 )

Substrate cone.(M) LiX(M) time(h) % reacted

fX KJ^

fx k^

F\

F\

Cl

^ s ^

Br

^ *J Cl

y-jT

L

DT

Λ

Λ

-Cl

-Br

1.1x10"2

9.3x10"3

6.1x10"3

5.0x10"3

-■

LiBr(0.18)

Lil (0.08)

-

LiBr(0.08)

Lil (0.08)

-

LiBr(0.18)

_

LiBr(0.18)

20

2

4.5

2

26

15

5

27

10

6

46

13

68

32

Polarity differences are also important in the course of photolytic processes as in the case of the dibromovinyl pyrethroid decamethryn where racemisations occur at different positions apparently depending on the polarity of the solvent (22).

Although examples are rare in the literature, it can be safely assu-med that aqueous solutions would favour charge transfer processes since the intermediate charged species could be readily stabilised by solvation. Charge transfer mechanisms are shown to be of importance in environmental photoreactions as in the electron transfer sensiti-sed photodegradation of DDT reported by Miller and Narang (12). These


findings were used by Parmar et al. in attempts to prepare photode-gradable formulations of DDT by incorporation of diphenylamine or di-ethyl aniline in the formulation (23). Diphenylamine also sensitises the photolysis of the ubiquitous pollutant hexachlorobenzene while benzophenone was found inactive. This suggests involvement of an elec-tron transfer process (24).

Although the structure of methoxychlor resembles that of DDT, the former compound is photodegraded much more rapidly both in organic solvents and in water (25). The primary photochemical process for methoxychlor appears to involve the initial formation of a free-ra-dical intermediate - scavenged by a thiol reagent - which yields the DDE analogue DMDE (Fig. 3). The thorough study of the photochemistry of methoxychlor by Zepp et al. (25) also showed that the photolysis rate was more rapid in several natural water samples than in distil-led water or when photolyses were performed in the presence of "hu-mic acid", Table 2. This substance, or complex of substances, have an absorption spectrum similar to that of materials dissolved in natural waters and the results indicate photodegradation routes of methoxy-chlor other than by direct absorption of light by the pesticide mole-cules.

hexane, 02

1CH3 CH3i Η 3 ( Γ ^ l ; ^ O C H 3

CCl2

Fig. 3. Major photodegradation pathways of methoxy-chlor (from Ref. 25).

2J0 G. Sundstrom and L. Octavio Ruzo

TABLE 2 Half-lives for photodecomposition of methoxychlor (40 ppb) under sunlight in various river waters (from Ref. 25)

Water half-life (h)

Distilled water, pH 6.3 300

Suwannee river, pH 4.7 2.2

Tombigbee river, pH 7.6 5.4

Alabama river, pH 7.7 2.9

Withlacoochee river, pH 8.2 no reaction in 2h

South Georgia stream, pH 7.2 "

20 ppm "humic acid" in dist water, pH 5.2 7.3

The work by Zepp and co-workers on methoxychlor (25) also may use as an example of different product distribution patterns upon photolysis in aqueous medium as compared to photolysis in hydrocarbon solvents (Fig. 3).

The influence of natural waters on photochemical reactions has also been investigated by Ross and Crosby who showed that the photochemi-cal epoxidation of aldrin to dieldrin can be enhanced in sterilised "paddy water" (agricultural water) probably by the presence of natu-rally occurring photochemical oxidants (26). Oxidative degradation of compounds not actually absorbing light directly is thus possible under certain circumstances in natural waters. This route of degrada-tion should not be readily distinguished from e.g. sensitising phe-nomena in complex natural waters.

In our laboratories the behaviour of several groups of polyhalogena-ted aromatic compounds, such as chlorinated biphenyls, naphthalenes, terphenyls and brominated biphenyls, have been studied. Several of these groups are known as environmental pollutants, e.g. the poly-chlorinated biphenyls (PCB), which are chemically related to the chlorinated pesticides and are commonly detected together in the en-vironment. All these compounds are highly lipophilic and are there-fore generally found accumulated in lipids of biological samples or in sediments.

The products of photolysis, mechanism and kinetics of the photoche-mical breakdown of PCBs have been investigated in several solvents (27-30). The types of products obtained vary considerably from polar phenolic compounds which may be converted further to more innocuous substances, to photoproducts that are as persistent as the starting material. Dechlorination, which is the most common photoreaction of PCBs, preferrably involves the chlorine substituents in the 2- and 6-positions of the phenyl rings. By loss of chlorine atoms in these positions the planar structure of the excited biphenyl molecule will be more favoured (28) :

Photochemical Transformation of Pollutants in Water

:i ci

cKe >κ )=α· ci-

211

In aqueous suspensions or solution and in methanol it has been shown that some isomers of PCB also may form dibenzofuran structures (31--34). Chlorinated dibenzofurans are considered almost as toxic as the well-known chlorinated dibenzo-£-dioxins but both types of compounds seem to be readily photodegraded as shown by laboratory experiments (31,32,35,36) and experiments performed under environmental conditi-ons (see below).

Studies of the photoreactions of chloronaphthalenes in hydrocarbon solvents showed that mainly dechlorinated products and binaphthyls are formed (37-39). Since the triplet energy of chloronaphthalenes is lower than 60 Kcal/mol (40,41) and the C-Cl bond strength larger than 80 Kcal/mol it was considered unlikely that the C-Cl cleavage reac-tion would be so endothermic and a charge transfer process was invo-ked to replace the original idea of homolytic C-Cl bond cleavage. This postulate was supported by the observation that the quantum yield of reaction increased considerably upon addition of amines and certain dienes, as well by the fluorescence quenching observed with triethyl amine (37,39). In aqueous solution the main photoproducts of 1-chloronaphthalene were binaphthyls, 1-naphthol and hydroxylated dimers. In the presence of oxygen the formation of dimers was sup-pressed and the yield of naphthols increased. The results shown in Table 3 may illustrate the variety of processes that may occur in the photoreactions of a single type of compound by varying the medium and added substances.

TABLE 3 Photoproduct distribution and quantum yields of some halonaphthalenes (from Ref. 37)

Substrate solvent 0v dehaloge-nation

binaph-thyls

substi-tution

Cl a

Br

MeOH MeOH-02 MeOH-Et3N MeCN-H20 Cyclohexane

MeOH MeOH-PhCOPh MeCN-H20 Cyclohexane

MeOH MeOH-PhCOPh

0, 0. 0,

0, 0.

0, 0,

.005

.002

.1

.007

.007

.012

.014

74 76 95 1

88

58 2 2

72

32 28

MeOH 0.17 5.2

25 23

94 12

38 97 94 28

66 68

48

5

4 1 4

2 4

%


The chlorinated alkylbiphenyls (chloroalkylenes) (42), considered as possible replacement products for PCB undergo photoreactions similar to those of other haloaromatic compounds (43). In methanol dechlori-nation occurs but no methoxy substituted products were observed (43). Polychlorinated terphenyls (PCT), however, give considerable yields of hydroxylated and methoxylated products when photolysed in water--acetonitrile or methanol, respectively (44,45) . In addition, PCTs may give high yields of diphenylene or triphenylene compounds in some solvents (Fig. 4), provided the PCTs are chloro substituted in the positions ortho to the phenyl rings.

Fig. 4. Some photoreactions of polychlorinated ter-phenyls (from Ref. 45).

As with halonaphthalenes electron transfer processes were implied in the photolysis of brominated biphenyls (PBB) (46,47) since conside-rable rate enhancements were observed in the presence of amines, which also increased the proportions of photoreduced products. The prefe-rence for ortho-debrominations, analogous to the findings with PCBs, were related to the stability of the radical ion by mass spectromet-ric studies (46).

The photoreactions of the widely used herbicides of the chlorophe-noxy acid type are well documented (48-54). Special interest has re-cently been devoted to the question of possible photoformation of the highly toxic 2,3,7,8-tetrachlorodibenzo-rj-dioxin (TCDD) from 2,4,5--trichlorophenoxy acid derivatives or degradation products thereof. The photodecomposition of TCDD - itself a byproduct in 2,4,5-T pro-ducts has also been investigated (36,55-57).

In water chlorophenoxyacetic acids photodecompose by cleavage of the ether linkage as well as by substitution of the chlorine atoms with hydroxyl groups. Sensitisation with acetone or riboflavin has repor-ted to increase the reaction rate for the free 2,4,5-T acid (50). However, Zepp et al. found that acetone did not sensitise the photo-reaction of 2,4-dichlorophenoxy acid (2,4-D) while the photolysis in certain natural waters was twice as fast as in distilled water (53). The photoreactions of 2,4,5-T in water are depicted in Fig. 5 and analogous schemes can be drawn for the photoreactions of other chlo-


OH OH

t t \ COOH OH

■Cl ^ Ά

I Q

At a

OCH2C00H ' OH

IT - * - ■ * - POLYMER

Fig. 5. Proposed photodecomposition pathway for 2,4,5-T in aqueous solution (from Ref. 50).

rophenoxy acids.

The photodynamics of 2,4-D esters (methyl and 2-butoxyethyl) in sur-face waters have recently been elegantly described by Zepp and co-wor-kers (53). The results were analogous to those obtained with the free acid, i.e. substitution of chlorine for hydroxyl groups. However, at concentrations exceeding the solubilities of the esters in water (>300 ppm) the major photoproducts were the dehalogenated 2-chloro-phenoxy- and 4-chlorophenoxyacetic acid esters, also major products upon photolysis in hydrocarbon solvents. The quantum yields for pho-tolysis of 2,4-D esters under different conditions are given in Tab-le 4 (53). The lack of efficient sensitising by acetone indicates that photolysis of 2,4-D esters proceeds from the singlet state (53).

TABLE 4 Dissappe« 313 nm (from

Compound

Ref. arance 53)

quantum yields

solvent

of 2 4--D esters

0

at

A*1

ci-AJ H2C00CH2CH2OCH2CH2CH2CH3

Cl

ct &

OCH2COOCH3

water, pH 5.3 " , pH 6.6 " , pH 7.8

n-hexane " -acetone

n-hexadecane

water, pH 6.6 n-hexadecane

0.056 0.052 0.056 0.17 0.0080 0.17

0.031 0.13

214 G. SundStrom and L. Octavio Ruzo

The photochemical work on the chlorophenoxyacetic acids has not pro-vided evidence for formation of TCDD upon photolysis (50,54). It has, on the other hand, been indicated that the small amounts of TCDD which may be formed in the environment readily undergoes photochemical breakdown. For example, by the use of a highly sensitive mass spec-trometric method we could not detect the expected residues of TCDD -based on the amount originally present in the herbicide preparation -on leaves, only a few days after application when dehalogenated pho-toproducts of the herbicide were detected (5). Similar results were recently reported by Crosby and Wong (55).

Photolysis of pentachlorophenol in aqueous sodium hydroxide (56,58) or alcohol (57) solution gives small yields of heptachloro- and oc-tachlorodibenzo-£-dioxins. Further photolysis of these compounds could theoretically give rise to TCDD or similarly substituted com-pounds, but laboratory experiments suggest that any TCDD formed is photodecomposed more rapidly than it is generated (32). Photoreac-tions of lower chlorinated phenols have been reported not to yield any chlorinated dibenzo-p-dioxins (50,54,56,59). However, in some ca-ses the formation of 2-phenoxy phenols ("predioxins") have been re-ported and these compounds in turn do give dioxins upon photolysis (54,60), Fig. 6. The negative findings of chlorinated dioxins in the photoreactions of chlorophenols may therefore be a question of sen-sitivity of the analytical procedure.

Munakata and Kuwahara found a number of complex dimeric and trimeric photoproducts upon photolysis of sodium pentachlorophenate in water solution under sunlight (61). Most of these compounds contained qui-nonoid structures with ether linkages and phenol groups.

Aromatic amines are important as such or in derivatised form (amides, ureas, carbamates) as herbicides. The photochemical reactions of aro-matic amines usually result in N-dealkylations and various types of rearrangements of the amine substituents. Photolysis of N-(tf-trichlo-romethyl-4-methoxybenzyl)-4-methoxyaniline in aqueous solution yields the dealkylated methoxyaniline hydrochloride and the rearranged and hydrolysed product N,2-dianisyl-2-oxoacetamide among several other compounds, Fig. 7 (62).

ΎΥΎΥ — Υ Ϊ Ύ Υ 1

cr^TOHcn^a c r ^ ^ o ^ ^ a

ci a

Fig. 6. Photochemical formation of chlorinated di-oxins from 2-phenoxyphenols (from Refs. 54,60).


vf-

CH3H2O

damp ether

H

C H ^ c r ^ A ^Λκΐ H2O

Fig. 7. Photoreactions of N- (0t-trichloromethyl-£--methoxybenzyl)-£-methoxyaniline in aqueous solu-

tions (from Ref. 62).

Basalin (N-(2-chloroethyl)-2,6-dinitro-N-propyl-4-trifluoromethylben-zeneamine) upon photolysis with sunlight in aqueous solution yields a number of products, many of which contain the benzimidazole struc-ture formed by reduction of the nitro groups followed by ring closure reaction of the amine substituent (Fig. 8) (63) . In laboratory experi-ments additional photoproducts were formed and identified (63). Ana-logous photochemical behaviour is shown by structurally related her-bicides such as trifluralin (<x,oc,oc-trl£luoro-2, 6-dinitro-N,N-dipro-pyl-4-toluidine) (64) and dinitramine (N^,N^-diethyl-2,4-dinitro-6--trifluoromethyl-m-phenylenediamine) (65).

Picloram (4-amino-3,5,6-trichloropicolinic acid) in aqueous solution photodecompose on both direct irradiation (300-380 nm)(66) and upon sensitisation (67) to yield a variety of unidentified products and chloride ion. Both free radical and ionic mechanisms have been sug-gested for the photoreactions (66) which presumably involve oxidation of the pyridine ring as a major photoprocess in analogy to the photo-oxidation of diquat (68).

The triazine herbicides photolyse in water to give mainly products formed by substitution of halogen or methylthio substituents with hydroxy groups (Fig. 9) (69-72). The rate of halogen replacement is dependent on the electron shell expansion capabilities of the halo-gen atoms (I>Br>Cl>F) (72).

In the photolysis of substituted phenylurea herbicides at environmen-tal wavelengths C-N bonds are cleaved and halogen atoms substituted for hydroxy groups (73-77). For example Rosen and co-workers have shown that monuron (3-(4-chlorophenyl)-1,1-dimethylurea) upon irra-


♦ W 2 ♦ V

C^NJLNOZ

% CH3CH2CH^N H

>%. 02NJWNO2 + 02NJCN02

Fig. 8. Photoproducts from natural sunlight irradi-ation of basalin in aqueous solution (from Ref. 63).

Cl CH Ν Λ Ν

Η2θ ψ Ν^Ν

RHN^N^NHR' R H N A N * S H R '

SCH3 H Ν Λ Ν Η2θ Ν^·Ν

RHN^N^HR' orMeOH RHN^N*S*fl'

Fig. 9. Major photoproducts of chloro- and methyl-thio-triazine herbicides (from Ref. 69).

diation with sunlight give substitution of the chlorine atom for a hydroxy group, while linuron (3-(3,4-dichlorophenyl)-1-methoxy-1--methylurea) is converted to the 4-hydroxy analogue as well as N-de-alkylation compounds (74) . Additional types of photoproducts of urea herbicides have been described by Kotzias and co-workers (77) and by Crosby et al. (75).

When a carbonyl group exists in the proximity to a proton, intramole-cular hydrogen abstraction followed by bond cleavage can occur, re-rembling the Norrish type II cleavage of ketones. Thus phenyl benzoyl ureas undergo cleavage to isocyanates and amides (76). Since the re-action also occur in the solid phase and in poor hydrogen donating solvents, the observed products must arise by intramolecular reac-tions, Fig. 10 (76).


hv » H o

»ΟΪΚ3 1

cr0

iMeOH

Ϊ hv

NHCOCH3 hV ^ γ Ν Η ί OCH3

Fig. 10. Photoreactions of 1-(4-chlorophenyl)-3-(2,6-dichlorobenzoyl)urea (from Ref. 76).

The photoreactions of phenylcarbamate insecticides were recently re-viewed by Silk et al. (78). These compounds are generally hydrolysed to phenols or rearranged to amides via photo-Fries reactions. A gene-ral photoreaction scheme for carbamates is given in Fig. 11. Zectran (4-dimethylamino-3,5-xylyl-N-methyl carbamate) yields upon photoly-sis in methanol solution N-demethylated and rearranged products in which the original 4-dimethylamino group has been substituted (79). Ivie and Casida have shown that aqueous suspensions of spinach chlo-roplasts are effective in sensitising the photodecomposition of the carbamates zectran and mesurol (4-methylthio-3,5-xylyl-N-methyl car-bamate) (11).

OCNHCH3 OH

CH3HNjj

Fig. 11. General photoreaction pathways for carba-mate insecticides (from Ref. 78).


The nitrodiphenyl ether herbicides photolyse in water forming phenols as is the case with e.g. fluorodifen (4-nitrophenyl-oe,0c,<x-trifluoro-2-nitro-^-tolyl ether) (80). Nitrofen (2,4-dichlorophenyl 4"-nitro-phenyl ether) yields the same types of products and since the phenols are formed in the same yield in both the absence and presence of oxy-gen a triplet excited state may be discounted and the reactions are believed to proceed by photonucleophilic substitutions (Fig. 12)(81). Other prominent reactions of the nitrodiphenyl ethers include reduc-tion of the nitro groups and formation of azo compounds (81,82).

chQ-0H -— a-^-oH

r JpH · H0-Q-0H -*- H0-Q-OH POLYMER

-Ο~Ν02^Η0^~)-Ν02

Fig. 12. Major photolysis pathways of nitrofen (from Ref. 82).

In organic hydroxylic solvents the simple chlorinated diphenyl ethers undergo reductive dechlorination reactions but yield no detectable amounts of phenols formed by cleavage of the ether linkage (83-85) . However, 2-chloro substituted diphenyl ethers give chlorinated di-benzofurans as photoproducts in several solvents (83-85) and upon sensitisation with acetone the yield of these toxic compounds increa-ses significantly (85). Photolysis of 2-chloro substituted diphenyl ethers in acetone can actually be used for the synthesis of pure di-benzofurans in nearly quantitative yields (86).

In the field of photochemistry of organophosphate insecticides two toxicologically interesting studies should finally be pointed out. Grunwell and Erickson found that parathion upon photolysis in aqueo-us tetrahydrofuran or ethanol, among other products, yielded 0,0,S-triethylphosphate (87). Since 0,0,S-trimethylthiophosphate has shown a synergistic toxicity with malathion the possibility of a similar effect between parathion and 0,0,S-triethylthiophosphate was sugges-ted in connection with poisonings among farm workers sometimes rela-tively long periods after spraying when the parathion itself had reached "safe" levels.


Sanborn et al. found that the desbromo photoproduct of leptophos (0-(4-bromo-2,5-dichlorophenyl) O-methyl phenylphosphonothionate) was a more effective neurotoxic agent than the parent compound and that this might explain why sheep and cattle fed leptophos-treated silage showed signs of intoxication quicker when exposed to sunlight (88).

ACKNOWLEDGEMENTS

Grants to one of us (G.S.) from the National Cancer Institute, USA, via the Organizing Committee of the 2nd International Symposium on Aquatic Pollutants and from the Research Committee at the National Swedish Environment Protection Board are gratefully acknowledged.

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( 7 1 )

( 7 2 )

( 7 3 )

( 7 4 )

( 7 5 )

( 7 6 )

( 7 7 )

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B.E

L.O

J . D

J . D

D.G

L.O

D.

P . J


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Oxidation of Organic Compounds in Aquatic Systems: The Free Radical Oxidation of Cumene

T. MILL, H. RICHARDSON andD. G. HENDRY

Physical Organic Chemistry Group, SRI International, Menlo Park, California

o The free radical oxidation of dilute aqueous solutions of cumene at 50 with a free radical initiator gives a complex mixture of products including acetophenone, cumyl alcohol, cumyl hydroperoxide and p-isopropyl phenols. The kinetics and mechanism of this oxidation have been investigated in pure water and the results applied to photooxidations of cumene in natural water in sunlight. Oxidation of cumene by a free radical process appears to be important in some aquatic systems.

INTRODUCTION

In 1975 the worldwide production of organic chemicals, excluding fuels and polymers, amounted to almost 408 billion pounds (1). The enormous potential for extensive and persistent pollution of water, air, and soil represented by these chemicals has led to a variety of actions to minimize environmental dispersion of many chemicals by controlling their use where persistence and biological effects are probable. Once these chemicals are dispersed into the environment, however, a variety of processes can effect the transformation of chemicals to less persistent and harmful forms. We believe that the key to designing and implementing effective measures to minimize environmental exposure of chemicals lies in detailed understanding of the environmental processes that control the concentrations and locations of these chemicals in water, soil, and air. Some processes such as hydrolysis are well understood (2); others, such as oxidation, are believed to be important but remain to be evaluated quantitatively.

This paper summarizes our present understanding of the chemical oxidation processes that could be important in water and describes in detail the use of cumene (isopropyl-benzene) as a probe for determining the kinds of oxidizing species that may be present in aquatic systems.

223

224 T. Mill, H. Richardson and D. G. Hendry

ENVIRONMENTAL OXIDATION PROCESSES

Free radical oxidation processes are undoubtedly important in certain natural waters. These processes involve combinations of elementary steps in which peroxy (RO #) or alkoxy (R0#) radicals abstract a hydrogen atom or add to a double bond in the substrate (RH); for example,

RO · + RH ► RO H + R · (1) 2 2

RO· + RH ► ROH + R· . (2)

In an aquatic system that may contain several kinds of RO · and RO* radicals, the total rate of oxidation of RH is the sum of the rates of individual rate steps (1) and (2) for each kind of radical. That is,

d[RH]/dt = k [RO.'il[RH] + k [RO·] [RH] , (3) 1 2 2

where it is assumed that reactions of all RO · or RO* radicals have about the same rate constants k or k . The half life t^ of a chemical reacting with both RO ' and RO* radicals is simply 2

t, = In 2/(k [RO ·] + k[RO-l), (4) 2 1 2 2

and since values of k and k are known for many kinds of chemicals (3), the problem of estimating t* reduces to one of estimating the values of [RO ·] and [RO·] that are likely to be found in the environment.

Table 1 summarizes values of t* for several classes of chemicals in reaction with RO · and RO· radicals. Values of t*, in Table 1 were calculated with the assumption that in aquatic systems, [RO ·] = 10 and [RO·-] = 1 0 M. Although neither of these values has been verified experimentally, they do provide the basis for evaluating the possible importance of these processes in qquatic systems. Half-lives are very long for most types of organic chemicals in reaction with RO · at the assumed concentration, but they are relatively short for reaction with RO·, even though the latter radical is at lower concentration.

Triplet diradicals such as (C H ) C-0 have reactivities similar to those of RO and could be important initiators of free radical oxidation in sunlight (4). Moreover, the products of oxidation such as ketones and hydroperoxides are sources of additional free radicals when exposed to sunlight.

Singlet oxygen ( 0 ) is a powerful but selective oxidizer formed usually by photo-sensitized processes:

1 hv 1 * Sensitizer ( sen) ► sen (5)

1 * 3 * sen ► sen (6)

3 * 3 1 1 sen + Ο Λ ► sen + On (7)

2 2

Oxidation of Organic Compounds 225

A typical reaction of 0 is concerted addition to double bonds to form either an 2

allylic hydroperoxide or a dioxetane.

1Ö + }:=CH-CH2- »-^-CH-CH- or fc-C-CH - (8)

00H 0-0

1 3 These reactions compete with conversion of 0 back to 0 .

2 2

(9) \> — 2

The rate of reaction with 0 is: 2

2

Rate = k [RH][ 0 ] . 8 2 ss

where [ 0 3 = km(k [RH] + k ). 2 SS 8 9

TABLE 1 Oxidation of Organic Compounds in Aquatic Environment b

Class

Alkane

Olefin

Benzyl

c Aldehyde

Alcohol

Chlorocarbon -CHC1-

Oxidizing Species

RO · 2

RO·

RO · 2

RO·

RO # 2

RO·

RO · 2

RO·

RO · 2

RO'

RO·

k2 (1/mol/sec)

0.005

4 13-10

0.09

4 24-10

0.1

4 11-10

3 3-10

7 10

0.01

4 15-10

4 0.5-10

Half-life, t, (days) *

5 2-10

0.5

2 9-10

0.3

8-103

0.5

0.4

0.1

4 9-10

0.5

15.0

Assumes [RO · ] = 10 M, [RO ' ] = 10 M. b 2 o

From R e f e r e n c e 3 ; m e a s u r e d a t 30 C. C F o r t h e r e a c t i o n RC0(0 · ) + RCHO.


Recently, Zepp et al (5) have reported that the concentration of 0 in some natural waters is as high as 2x 10 M. At this concentration, certain organic sulfides, furans, and amino acids would react rapidly enough to be oxidized at significant rates, but other types of chemicals would be unaffected· However, small amounts of peroxides formed from O in reaction (8) could be a source of free radicals that initiate other oxidation processes.

Figure 1 summarizes some of the complex interrelationships that may exist among peroxides, free radicals, excited species, and sunlight, which can drive a cyclic oxidation scheme. Our ability to predict the rates and pathways of oxidation reactions in aquatic systems is seriously limited by lack of information about the identity and concentrations of important oxidizers in aquatic systems. We have initiated research to determine the rates at which selected organic compounds oxidize under environmental conditions and the products resulting from these oxidations. This information should allow us to determine the kinds and concentrations of intermediates involved in aquatic oxidation processes. Cumene was selected initially for detailed study for several reasons:

(1) The rates and products of oxidation of cumene by free radicals such as RO · or RO* are well understood in nonaqueous systems (6).

(2) Primary products of oxidation are relatively stable and readily analyzed at low concentrations in aqueous solutions.

(3) Cumene is not particularly susceptible to enzymatic oxidation or to oxidation by singlet oxygen.

OXIDATION OF CUMENE IN WATER

The free radical oxidation of cumene in nonaqueous solution gives varying proportions of cumene hydroperoxide, cumyl alcohol, and acetophenone by the following steps (6):

k. I n i t i a t o r ► 2eR0 · (10)

2

CuH + RO · ► Cu· + RO H (11)

f a s t , v Cu· + 0 - *-CuO ' (12) 2 2

k CuO · + CuH £-CuO H + Cu· (13)

2 2 ak

2CuO · ►Cu 0 + 0 ( t e r m i n a t i o n ) (14) 2 2 2 2

v ( l - a ) k x ^2CuO? + 0 (15) k 2

CuO· +CuH—2-CuOH + Cu· (16)

k d CuO·—►C H COCH + CH · (17)

6 5 5 3

^R

0oH

SO

UR

CE

S

AN

D

RA

DIC

AL

C

HA

INS

D

O

X a.

o

H o

o I o a.

Occurs Naturally

Fig. 1.

Intermediates and pathways for environmental oxidation.


High concentrations of cumene give longer chain lengths and CuO H as the major product; the kinetic relationships are:

\ d[CuH]/dt = (R /2ak ) k [CuH]

i x p \

d[0 ]/dt = (R /2ak ) k [CuH] + R /2a 2 i x p i

where R = 2ak [Initiator], i i

At low concentrations of cumene, reactions (13) and (16) become slower than reactions (14), (15), and (17), and the major products are CuOH and C H COCH ; CH * becomes

6 *$. 3 3 an important radical in the termination processes since reaction (19) supplants the slower terminating reaction (14):

CH * + 0 — ► CH 0 · (18) 3 2 3 2

k CH 0 · + CuO •—►CuOH + CH 0 + 0 . (19) 3 2 2 2 2

Under these conditions, the kinetic relationships become complex.

o We have examined the radical-initiated oxidation of cumene at 50 C in water using a water-soluble azo initiator azobis(4-cyanovaleric acid) (AA) to provide a steady source of RO · radicals. Initial concentrations of cumene ranged from. 10 to Ιθ" Μ. Table 2 summarizes the results of several experiments that were analyzed for cumene and its oxidation products, usually after 100 hr of reaction time at o

50 C. Product analyses summarized in Table 2 show that in addition to the products expected from radical oxidation—such as hydroperoxide, alcohol, and ketone, we also found significant yields of the isomeric isopropylphenols. These products have not been reported before in the oxidation of cumene and require some special explanation. The absence of dicumylperoxide also requires comment. Coupling products of cumene radicals and initiator radicals probably are also present in these mixtures; they have been detected but not quantified as yet. In the absence of the radical initiator, no oxidation products were detected.

Kinetic Analysis of Cumene Oxidations

The very low concentrations of cumene in water containing high concentrations of radical initiator creates a special set of conditions not usually encountered in liquid phase oxidations. The initial rate of production of radicals in water at 50° from 10" 4 M AA (k = 2.10-6 M-* sec"1) is as follows:

d

r Ί -10 _1 R = 2ek [AA] = 2.4 x 10 M sec i d

However, the average rate of oxidation of cumene in experiments 46, 86, and 60 (Table 2) is only 2 x 10" M sec" ; thus, the average kinetic chain length is less than 0.1. It is useful to view these oxidation systems as consisting mostly of radical-radical interactions with only a minor fraction of the radical (< 10%) involved in H-atom transfer from cumene. Thus, the rate of oxidation of cumene

TABLE 2 Oxidation of Cumene in Pure Water at 50°C

Exp

eri

No.

1 2 3 4 5 6 7 8 9a

, .men

t

b,c

(c

on

tro

l)

Tim

e (h

r)

64

62

64

100 96

96

96

96

72

Rea

cta

nts

[CuH

]*

20

36

43

10

d 1

3.5

d 14

4

e 16

0

205

20

9.6

±

5.7

100

100

100 99

100

100

100

100

0

[Cu

H] f

13

23

27 1.0

6.7

138

143

189

20

9.3

±

7.6

Co

nce

ntr

ati

on

s

TC

uOH

]

0.9

3

0.9

7

1.1

2

0.1

1

1.3

3

4.4

4.2

3.6

0

[Cu

0 2H

]

0.9

8

1.3

1.3

0.1

2

8.2

11

.8

4.8

6.2

0

[C in

M

olar

6H

5C

0Me]

0.0

86

0.7

8

0.1

0

0.0

15

0.1

2

0.3

1

0.1

4

0.4

6

0

x 1

0e

Pro

du

cts

p-i

-c

■P

rC6H

40H

2 2 2 0.5

1.5

3.8

4

7.6

4.7

0

o-i

-PrC

6H

40

HC

.

0.5

0.4

0.4

0.1

0.2

5

0.8

1

1.2

1.1

0

[ACu

H]

7.0

13

16 9.0

6.8

6.0

e (3

3)

16

0

>H

XI

10

11

Mse

c"-1

3.0

5

.8

6.9

2.5

1.9

1.7

9 4.4

0

8 H·

Pi

rt

H·

O o H-i

o OQ

fa

H·

O

O

O •8

O c en

a Purified cumene containing no detectable products.

Azobis(4-cyanovaleric acid).

Identification of phenols in first four experiments is tentative.

Experiment carried out with oxygen instead of air

e Estimated from 2 x (sum of products) plus [CuH]

[CuH

]*


depends only on the concentrations of initiator peroxy and alkoxy radicals formed

by a reaction sequence similar to that of CuO · radicals:

°2 In ► 2eR02· (10)

RO · + CuH ► RO H + Cu· (11) 2 2

(l-a)k 2RO ' £► ROOR + 0 (termination) (20)

2 2

^ 2R0· + 0 2 (21)

RO· + CuH ► ROH + Cu· (22)

RO· > Cleavage product + Me# (23)

Me· + 0 ► MeO · (18) 2 2

RO · + MeO · ► ROH + CH20 + 0 (termination). (24)

A more detailed analysis of the fate of RO· radicals in this system suggests that

they will always cleave to give NCC(0)CH2CH2COOH and Me· radicals rather than

abstract. Under these conditions

1/2 [RO ·] = (R./2k ) ,

^ i x

and

1/2 d[CuH]/dt = k [CuH](R /2k ) . (25)

p i x

This scheme also accounts for the fact that no Cu O was detected from reaction

(14), simce most CuO . radicals will interact wi • and MeO · radicals as * n 2 2 2 follows:

CuO · + RO · ► ROOCu + O (26)

CuO · + MeO · ► CuOH + CH O +·θ . (27) 2 2ι 2 2

Thus the CuOH found is formed in reaction (27), not in reaction (16). Aceto-

phenone, which can only arise from cleavage of CuO·, provides an accurate measure

of the total CuO· formed; Table 2 shows that C H COMe is only 1 to 6% of the

alcohol and hydroperoxide. Over 90% of the alcoRol must be formed by radical-

radical interactions or other pathways not involving CuO.


From equation (25) we can calculate the rate constant for reaction (11), which

entails H-atom transfer from cumene to RO ·: 2

1/2 k = [d(CuH)/dt]/(R /2ak ) [CuH] p i x av

5 _1 _1 where ak = 1 x 10 M sec and values for other terms are given in Table 2.

x

Values of k were averaged for experiments in Table 2; for the 62-64 hr experi-

ments, k ~ 300 M" sec" , and for the 100 hr experiments, k ~ 170 M~ sec" . For

comparison, in the reaction

CuO · + CuH ► CuO H + Cu* , 2 2

_1 -1 the value of k is 0.47 M sec . The large difference between values calculated

here and for reaction in pure cumene may be partly accounted for by the higher

reactivity of RO · than CuO · in H-atom transfer. However, the discrepancy is

still large and may indicate that some RO· radicals, which are much more reactive

than RO · , may contribute to the total rate of reaction. 2

The most surprising finding in this study is the formation, in amounts equal to

the alcohol and hydroperoxide, of o-, and p-isopropyl phenols. The ratio of

isomers found here and the apparent absence of the meta-isomer are at least in

accord with the idea that radical-radical couplings, similar to those found for

phenolic oxidation systems, are responsible for the formation of these phenols:

CuH + R0 · - Cu· + RO H 2 2

Cu· + 0 < * CuO · 2 2

<>-, ^.Q^.»,— ( \ Q ^

°=Q<

Me

Me + RO·

Me

Me RQH

Qualitatively this scheme is in accord with the observation that in experiment 6,

where pure oxygen was used, more hydroperoxide and less phenols were found than in

a similar experiment, 7, where air was used. Use of more oxygen should favor

formation of more CuO · at the expense of the cumyl radical and its isopropylidene

232 T. Mill, Η. Richardson and D. G. Hendry

cyclohexadiene resonance form. However, the effect is small and additional exper-iments are needed to verify this proposed mechanism.

Another plausible mechanism for phenol formation entails initial electron transfer from Cu* to 0 to form

2 to form both alcohol and phenols

transfer from Cu* to 0 to form the carbonium ion. which then reacts with water 2,

i -l Λ Cu· + 0

2 +

Cu + H 0 2

+ ^ Cu + · 0

2 +

' CuUH + H

Cu+—-^ H^/ V ^ + HO ► H0-<^ \ <

Intervention of an ionic process in water may be energetically favorable and would account for the absence of phenols when cumene is oxidized in nonpolar solvents.

Oxidation of Cumene in Natural Water

Several experiments have been carried out using cumene to probe the oxidation processes occurring in natural water samples taken from a ear San Jose, California. The creek water, which is slightly yellow, alkaline (pH ~ 8), and odorous, shows a significant polar material content by hplc with trace amounts of (probably) volatile organics, none which are cumene or its oxidation products. Water samples were filter-sterilized before use by passage through a 0.22- /im filter. Experiments were conducted at 50° in the dark and at ambient temperatures in sunlight, using added cumene with and without added free radical initiator azobis(methylisobutyrate) (MAB). Table 3 summarizes the experi-mental resu1ts and Fiyure 2 illustrates the hplc traces for these reaction mixtures.

When water samples alone were irradiated in July sunlight for 20 calendar days, no products with elution times similar to those from cumene were found by hplc. How-ever use of MAB, a nonionic soluble azo initiator, did give oxidation products with cumene in pure and natural water at 50°C, including acetophenene, phenols, and cumyl alcohol, but these occurred in different proportions than those observed with AA in pure water. Much more acetophenone was found in these experiments than in earlier experiments, along with a variety of (probably) coupling products of cumyl peroxy and initiator peroxy radicals; these products were not visible in hplc when AA was used as the initiator.

Cumene dissolved in natural water and exposed to July sunlight for 20 days formed acetophenone and isopropyl phenols in nearly equal amounts and at about 20% of the rate of formation of these same products using MAB or AA at 50° in pure water. Sunlight irradiation of cumene in pure water for the same period of time gave barely detectable traces of acetophenone.

We cannot determine from these analyses whether additional cumene radical coupling products are associated with natural materials, because a material balance on

TABLE 3

Oxidation of Cumene in Natural Water

Experiment

No.

20

21

22

23

24

25

26

Time

(hr)

240

240

96

240

129

96

96

Reactants

[CuH] Q

0

200

200

200

200

200

200

[Initiator]c'd

sunlight 30°

0,50°

0,25°

sun, 30°

sun, 30°

100

AA, 50°

100

MAB, 50°

[CuH]f

0

200

200

200

~200

-186

[CUOH]

0

0

0

trace

0

4

[CuOH]

0

0

0

< 0.1

0

< 0.1

Products

[C H COMe]

6_J>

0

0

0

0.97

trace

4

i-PrC H OH

6 4

0

0

0

0.8

0

4

Others

0

0 —

—

—

—

8 peaks

o

Γΐ

o

O

Hi

O

H·

O O O 1

o

c

Water from Coyote Creek, San Jose, California.

Hours of sunlight or heating.

AA = azobis(4-cyanovaleric

acid); MAB = azobis(methyl isobutyrate),

Temperature in sunlight experiment varied from 15-35°C.


i-PrC6H4OH

Photooxidized CuH in Water

CuH in Water in Dark

Water Background

10 15 20 25

RETENTION TIME (minutes)

30

Fig . 2 . High pressure l i qu id chromatograms for photooxidat ion mixtures from cumene in eutrophic water (Coyote Creek) 0-100% a c e t o n i t r i l e in H2O; l i n e a r g r ad i en t , 2 ml/min; 254 nm, 0.04 absorbance u n i t s fu l l s c a l e .


cumene was not obtainable. Nor have we determined whether products such as

hydroperoxide and phenols may have been formed and then photolysed. It is

possible that the high proportion of acetophenone observed in these experiments

arise from photolytic or metal ion catalysed cleavage of hydroperoxide to form

CuO· radical, which then cleaves to acetophenone:

CuOOH ► CuO* + HO*

CuO· ► C H COMe + Me· 6 5

The clear implication of these results is that free radical processes do take

place at measurable rates in natural waters and that peroxy radicals are formed

in some stage of the process.

CONCLUSIONS

The oxidation of cumene at low concentrations in water provides a useful and

specific chemical probe for the presence of free radicals in natural waters

irradiated by sun light. The detailed mechanism of oxidation under these condi-

tions to give oxidation both at the tertiary C-H bond and in the aromatic ring

is different from that observed at higher concentrations in nonpolar solvents.

Possibly an electron transfer process becomes important in water and leads to

formation of phenols.

The experiments in natural waters irradiated by sunlight gave about the same order

of magnitude of products as that found in distilled water using a free radical

initiator at 50°. However, the composition again is different in the two cases,

with much more acetophenone formed in natural water in sunlight than in distilled

water at 50°.

A likely source of radicals in sunlight is carbonyl sensitizers that can form

triplet diradicals with reactivity similar to that of RO· radicals.

Additional experiments are now under way to characterize the importance of oxida-

tion reactions in other natural waters and to more fully elucidate the mechanisms

of oxidation of cumene under these conditions.

ACKNOWLED GMENTS

This work was supported by the National Science Foundation-RANN under grant no.

ENV76-11153.

REFERENCES

1. U„S. Tariff Commission Data for 1975.

2. W. R. Mabey and T. Mill, J. Phys. and Chem. Ref Data. 7, 000 (1978).

3. D. G. Hendry, T. Mill, L. Piszkiewicz, J. A. Howard and H. K. Eigenmann, J.

Phys. and Chem. Ref Data, 3, 937 (1974).


4. J. A. Howard, "Advances in Free-Radical Chemistry" Academic Press, New York, p. 49 (1972).

5. R. G. Zepp, N. L. Wolfe, G. L. Baughman, and R. C. Hollis, Nature, to be published.

6. D. G. Hendry, J. Am. Chem. Soc. 89, 5433 (1967).

Prediction of Photochemical Transformation of Pollutants in the Aquatic Environment

RICHARD G. ZEPP and GEORGE L. BAUGHMAN

Environmental Research Laboratory U.S. Environmental Protection Agency, Athens, Georgia, U.S.A.

ABSTRACT

Mathematical models that predict concentration-time profiles can be used to assess potential exposure of aquatic organisms to toxic pollutants. Such models integrate quantitative descriptions of individual transport and transformation processes that affect pollu-tant behavior in the environment. Discussion in this paper focuses on methods for prediction of one important transformation process, photolysis by the action of sunlight.

Two general classes of photochemical transformation are direct photolysis, initiated by direct absorption of light by the pollutant, and indirect or sensitized photolysis involving light absorption by natural "photosensitizers." At concentrations normally encountered in natural waters, direct photolysis is described by first order rate expressions. Equations that employ solar irradiance, quantum yields, and absorption spectra of pollutants are used to predict photolysis rates as a function of time of day, season, location, and water depth. Sensitized photolysis can proceed by a variety of mechanisms including energy transfer, sensitizer-pollutant reaction, and oxidation involving singlet oxygen or free radicals. Recent studies have indicated that singlet oxygen, a species that rapidly oxidizes certain pollutants such as polycyclic aromatics, is generated photochemically in a variety of natural waters. Although qualitative evidence indicates that other types of sensitized photolysis are important, mechanisms and equations required for quantitative description of these processes are still poorly defined.

237

238 R. G. Zepp and G. L. Baughman

INTRODUCTION

The time-honored approach for deducing the environmental fate of a man-made chemical has been to study its behavior under field conditions. Usually, the approach has involved collecting and analyzing large numbers of samples from a portion of the environment, such as a river or field, that has been exposed to the chemical. To complement this approach, qualitative laboratory studies are used to explain the observed field behavior.

Carefully planned field studies still remain the best way for evaluating the behavior of chemicals in the environment. Public concern over the toxicity and adverse ecological effects of certain chemicals, however, has stimulated interest in predicting which chemicals might cause harm to susceptible organisms. Because few field data on these compounds may be available, initial decisions concerning their environmental safety will be based largely on laboratory data. To adequately assess the potential hazard of chemicals to aquatic organisms, toxicity data must be accompanied by some way of predicting the persistence and movement of the chemicals in aquatic ecosystems. Two laboratory approaches appear most promising. One, discussed elsewhere (Ref. 1), involves observing the distribution of chemicals when added to small model ecosystems. The other approach uses laboratory data concerning various environmental processes that determine the transport and transformation of chemicals in the aquatic environment. During the past few years, our research program has developed techniques for extrapolating quantitative data concerning microbial, hydrolytic, photochemical, and sorption processes from laboratory to environmental conditions (Ref. 2). We have used these techniques to assess the likely importance of various environmental pathways for degradation of chemicals (Ref. 3). Recently, Hill and coworkers (4) and Smith et al. (5) have shown how quantitative laboratory data concerning toxic chemicals can be integrated in mathematical models to predict concentration-time profiles in various environmental compartments. This paper is concerned with laboratory procedures for forecasting the photochemical transformation of chemicals in the aquatic environment.

Energy from the sun is the primary driving force for degradation of chemicals in the aquatic environment. It is well-recognized that solar energy stored in biological systems may eventually be used to transform natural and man-made chemicals. The chemical structures of synthetic chemicals, however, are often very different from those of natural substances. Thus, considerable periods of time may be required for acclimation and subsequent degradation of these chemicals to occur by biological pathways. However, other non-biological pathways for transformation are also available. Sunlight absorbed by the atmosphere and water bodies indirectly provides the energy for spontaneous, thermal reactions such as hydrolysis and oxidation (Ref. 6). Finally, the simplest pathway for degradation involves the direct conversion of absorbed solar energy into photochemical reaction. The most obvious evidence of photoreaction in aquatic environments is the widespread presence of phytoplankton and other light-dependent under-water plants. Much effort has been directed towards mathematically describing photosynthesis by freshwater and marine phytoplankton (Ref. 7), but most of this work is not directly applicable to predicting photochemical reaction of synthetic chemicals.

Prediction of Photochemical Transformation 239

Any photochemical reaction depends upon the dose of light received and the response of the irradiated system. Most of our discussion here centers upon prediction of the response of chemicals in those portions of aquatic environments that are exposed to sunlight. Later in this paper, we will discuss factors affecting the dose of light received by pollutants and outdoor verification of the predictions based on laboratory data.

PATHWAYS FOR PHOTOLYSIS IN AQUATIC ENVIRONMENTS

A number of competing photochemical pathways may determine the response of a chemical exposed to sunlight. For convenience, we have subdivided these various pathways into two general classes of photoreaction, direct photolysis and indirect or sensitized photolysis.

Direct Photolysis

Theory and equations. The simplest mechanism for photolysis involves direct absorption of light, which produces a molecule in an electronically excited state (eq. 1). The excited molecule either reacts (eq. 2) or decays by some mechanism back to its ground state (eq. 3).

P0 + Light > P* (1)

P* —^-> Products (2)

p* JLL> p0 (3)

Most studies of aquatic pollutants have involved elucidation of products derived from direct photolysis in non-absorbing solvents such as water, methanol, or hexane.

The mere passage of light through a system does not ensure direct photolysis. The light must be of wavelengths that correspond to electronic absorption by the chemical. Thus, it is entirely possible that a pollutant that reacts rapidly under sunlight could be unreactive when exposed to a lamp with footcandle intensity similar to that of sunlight. Units of luminosity such as footcandles refer to visible light, whereas many pollutants absorb only the ultraviolet (uv) portion of sunlight. On the other hand, chemicals such as DDT that photoreact rapidly when exposed to uv light from mercury or xenon lamps (even with filtering by Pyrex) can be very unreactive outdoors because of a lack of sunlight absorption (Ref. 8).

Clearly, the rate* of direct photolysis is related to the rate Ia at which light is absorbed by the pollutant. Under most natural conditions, the concentration of a pollutant [P] is very low and Ia is directly proportional to [P] (Ref. 9). The proportionality constant, ka, is expressed in units of reciprocal time and is called the

* The rate of photolysis is defined as the change in concentration per unit of time, -d[P]/dt.


specific sunlight absorption rate. The value of k can be calculated using solar irradiance incident on the surface of the water body, absorption coefficients of the pollutant, and attenuation coefficients of the water body (Ref. 9).

If all of the light absorbed by the pollutant were converted into photoreaction, the photolysis rate would simply equal Ia. In fact, usually only a fraction of the absorbed light causes chemical change. This fraction is called the quantum yield for reaction, φ. The rate expression for direct photolysis is thus:

- d [ P ] - ΙΛ φ = ka φ [Ρ] = k*[P] (4) dt a

To make the data easier to visualize, we also have expressed them in terms of half-lives inherent to direct photolysis, t :

u _0^93_ (5)

h k

The value of ka in a water body depends strongly on light screening by natural substances in the water. The screening effect will be discussed later in the paper. Near the surface of a water body, the pollutant is completely exposed to sunlight and ka becomes independent of the optical properties of the water body. The near-surface half-life, tf, can be computed by eq. (6):

to = 1-81 x 102° (6)

* Φ Σ ελ h

where ε^ is the molar absorptivity of the pollutant in 1 mole"1 cm"1 at wavelength λ, Ζχ is solar irradiance at a wavelength interval centered at λ in photons cm"2 sec"1, and the term Σε^Ζ^ indicates summation over the wavelengths of sunlight absorbed by the pollutant. Tables of Z values have been published elsewhere (Ref. 9). Molar absorptivities, or "extinction coefficients," can be easily obtained with modern spectrophotometers; mixtures of water and acetonitrile or methanol can be used as solvents for these determinations when solubility limitations in water are a problem. Zepp (10) has recently described procedures for measuring quantum yields for reaction of pollutants in very dilute aqueous solution.

Predicted direct photolysis rates. Zepp and Cline (9) have developed a computer program that is used to calculate solar spectral irradiance as a function of time-of-day, season, and latitude. The program uses solar irradiance, quantum yields, and eq. (4) to compute direct photo-lysis rates of pollutants in water bodies. Results of the computations for the widely distributed pollutant, benzo[a]pyrene, illustrate how direct photolysis varies with time and location (Fig. 1 and 2). Molar absorptivities and quantum yields obtained by Smith and coworkers (5) were used in these computations. Seasonal variations are predicted to be more pronounced in Europe (latitude 50°N) than in the mid-U.S. (latitude 40°N), but during the summer months, there is not much difference in half-life (Fig. 1). Changes in photolysis rates are caused by fluctuations in solar irradiance as well as by


variations in the period of daylight (Fig. 2). To facilitate comparison of photolysis to other processes that are not light-dependent, the data in Fig. 1 are integrated over the 24-hour day (Ref. 3). Data summarized in Table 1 indicate that many chemicals are rapidly transformed upon exposure to sunlight.

>-

^ . TS

LF

-LI

cc I

I . M -

\

\

\

\

\

\

\

^ , \

\

\

XE Lfl LO UP DE

10BI0 ■ITUD JOITU FR B

»TH

r ic , 5E ! 1ΠΥ .

BENZ 4 0 . 5 90 PIIRF NEAR

HflDP J. Θ0

UflTF SURF

YRENE

RCE

^

/

J //

60

/

/

/ /

7 /

Λ \

/ /

,H /

MONTH OF YEAR

Fig. 1 Annual variation of the near-surface half-life (t^) of benzo[a]pyrene at several Northern latitudes 4

o (/)

O

3 h

— 2 h

σ

Summer Solstice

0 3 0 0 0700 MOO 1500

LOCAL TIME 1700

Fig. 2 Diurnal variation in direct photolysis rate of benzo[a]pyrene computed for Amsterdam, The Netherlands


TABLE 1 Predicted Near-surface Half-lives, t£, for Direct Photolysis of Pesticides and Polycyclic Aromatic Hydro-carbons - Midsummer, Latitude 40°N a> b

Chemical

(1) N-Nitrosoatrazine (2) Trifluralin (3) DMDE (4) DMDE (5) DDE (6) Diphenylmercury (7) Phenylmercuric acetate (8) 2,4-D butoxyethyl ester (9) Carbaryl (10) 2,4-D butoxyethyl ester (11) Benzo(f)quinoline (12) Dibenzo(e,g)carbazole (13) Benzo(a)pyrene (14) Benz (a)anthracene (15) Carbazole (16) Dibenzothiophene (17) Quinoline (18) Methyl parathion (19) Parathion

(a) Half-life integrated over 24-hour day. (b) Data from Zepp and Cline (9) and Smith et al. (5)

Solvent

Water Water Water

Hexadecane Water Water Water

Hexadecane Water Water Water Water Water Water Water Water Water Water Water

Quantum Yield for Reaction

0.30 0.0020 0.30 0.20 0.30 0.27 0.25 0.17 0.0060 0.056 0.014 0.0033 0.00089 0.0033 0.0076 0.00050 0.00033 0.00017 0.00015

t°, da

0.0092 0.039 0.094 0.14 0.92 0.80 1.6 3.8 6.6 12.0 0.060 0.033 0.059 0.066 0.12 5.7 25.0 9.0 10.0

Influence of natural substances. As a first approximation in computing HTrect photolysis rates, we assume that the electronic absorption spectrum and quantum yield for reaction are the same in natural as in air-saturated pure water. How close is this assumed rate to the actual photolysis rate in a water body? In this section we discuss the likely effects of the dissolved and suspended matter in natural water upon direct photolysis.

Of course, if the pollutant rapidly undergoes a chemical reaction in a natural water that changes its chemical nature, then its photochem-istry will change. Examples of such chemicals are certain phenols, such as pentachlorophenol, that dissociate to phenoxide ions in basic natural water. Also, the photoreactivity of other ionizable chemicals can be changed by complexation with natural species. For example, methylmercuric chloride, itself non-photoreactive, may be transformed into photoreactive sulfur-bonded complexes in some aquatic environments (Ref. 11). Likewise, NTA and EDTA (Refs. 12,13,14) become photoactivated when they form complexes with metal ions, such as iron (III), that are commonly found in lakes and rivers. Thus, to adequately forecast the photochemical behavior of ionizable chemicals, spectra and quantum yields of environmentally significant dissociated or complexed forms must be obtained.

Sorption of chemicals onto sediments in aquatic environments also has several important effects upon photolysis rates. Hautala has concluded that sorption of pesticides onto soil surfaces markedly diminishes their quantum yields for reaction (Ref. 15). The reaction


conditions in his experiments were considerably different from those found in natural waters, however. McGinnes reported that sorption onto a clay suspended in water had little effect on the direct photolysis rates of benzo[a]pyrene and benz[a]anthracene (Ref. 16). Recently, Miller examined the direct photolysis of several pesticides in the presence of sediments obtained from rivers and ponds in the southern United States (Ref. 17). He concluded that the effects of suspended sediments on the photochemical response of aquatic pollutants are likely to be unimportant compared to the diminution of the dose of sunlight caused by the presence of sediments {vide infra).

Dissolved species in natural water may alter quantum yields for reaction through interactions with the excited state, P*. "Quenching" interactions return the electronically excited pollutant to its ground state (eq. 7).

P* + Q — ^ > P O (7)

The effect of quenchers upon the quantum yield for reaction is represented by the Stern-Volmer relation (eq 8).

-ί2- = 1 + *7 x0 [Q] (8)

where φ0 and φη are the quantum yields in the absence and presence of quencher, respectively; k7 is the rate constant for quenching; τ0 is the lifetime of P* in the absence of quencher; and [Q] is the concentration of quenchers (Ref. 18). Dissolved molecular oxygen is known to be an excellent quencher (Ref. 19), and because it is present at concentrations higher than other species in freshwater (Ref. 11), the quenching effect is adequately accounted for by simply conducting direct photolysis studies in air-saturated pure water. Excited state lifetimes of the pesticides, 2,4-D (Ref. 3) and methoxychlor (Ref. 8) are so short that their quantum yields should not be decreased by quenching in freshwaters. On the other hand, photoreactions of carbaryl (Ref. 6a) and probably polychlorinated biphenyls (PCB's) (Ref. 20) are partially quenched by oxygen in air-saturated water.

Interactions of P* with natural species may also involve chemical reaction (eq. 9).

P* + Y > Products (9)

Direct photoreactions of polycyclic aromatics with oxygen are ade-quately simulated in air saturated pure water. Photoreactions with other species may increase the quantum yield for reaction, and thus the photolysis rate, over that observed in pure water and also may change the reaction products. For some time it has been known that certain aromatic compounds are susceptible to photoinduced nucleophilic substitution reactions (Ref. 21). Crosby and coworkers have provided ample evidence that many pesticides can undergo photonucleophilic reactions in the presence of high nucleophile concentrations (Ref. 22). Ruzo and coworkers have demonstrated that PCB!s also undergo light-induced nucleophilic displacements (Ref. 20). Although nucleophilic species such as hydroxide, chloride, bromide, thiols, and sulfide are present in freshwater environments (Ref. 11), their concentrations are typically far too low for them to compete with nucleophilic displacements by water itself. Sea water, however,


may have sufficiently high concentrations of chloride (0.6 M) and bromide ( 10"3 M) to compete with water.

In summary, because excited state lifetimes are very short (usually less than 10 nanoseconds) and because concentrations of chemical species are very dilute in aquatic environments, it is reasonable to assume that quantum yields for reaction are usually the same in pure, air-saturated water and natural water. Exceptions to this general rule may occur with some chemicals. Dissociation or complexation can, of course, alter the chemical nature, and thus, the photoreactivity of ionizable synthetic chemicals.

Sensitized Photolysis

Photosensitizers are chemicals that, upon light absorption, cause chemical changes that do not occur in their absence. That photosensitized oxidations occur in many biological systems has been known for some time (Refs. 18,23), but it was not demonstrated until very recently that photosensitized transformation of pesticides can occur on solid surfaces such as plant leaves (Ref. 24) or in aqueous solution (Refs. 3,6a,8,25,26,27). Pioneering studies by Crosby and Ross (25) demonstrated that ethylene thiourea, a pesticide metabolite, underwent photosensitized oxidation in water from an agricultural field, although it is inert to sunlight in distilled water. Zepp and coworkers have found that photosensitized transformations of a variety of pesticides occur in natural water (Refs. 3,6a,8,26). For example, the photochemical transformation of the insecticide, parathion, is greatly accelerated in water obtained from the Okefenokee Swamp in southern Georgia, U.S.A., and in many other natural waters as well (Fig. 3) (Ref. 28). In this section, we discuss several mechanisms that can account for these photosensitized transformations.

0-P-(OC2H5)2

0 4 8 12 16 20

TIME, Hrs. 24 28

Fig. 3 Comparison of photolysis of parathion in distilled water and in a water sample obtained from the Aucilla River, Lamont, Florida


Energy transfer. One important mechanism involves the transfer of absoroed light energy from sensitizer to pollutant (eq. 10-14).

S + Light > XS*

i s* ^ 3 S*

3s* + p —L2__^ s + 3P*

k| *S* + 02 ~13 > S + 0 *

(10)

(11)

(12)

(13)

^ 1 I» (14)

Two systems of excited states are available to the sensitizer, the singlet, 1S*9 and triplet, 3S*, states. Light absorption initially puts the sensitizer into its excited singlet state, but energy transfer usually occurs from the longer-lived triplet state, 3S*. Under environmental conditions, most of the transferred energy goes to dissolved oxygen because the concentration of oxygen, an excellent energy acceptor, is typically several orders of magnitude greater than pollutant concentrations. The rate equation for a photosensitized reaction under environmental conditions is:

Rate = Ij φ15 φ 6 ί φκ (15)

where l| is the rate of light absorption by the sensitizer, (|>jg is the fraction of absorbed light that results in triplet formation, <j>R is the quantum yield for reaction of the pollutant, and φθ-£ is the fraction of triplet energy that is transferred from 3S* to the pollutant. The value of <J>et is:

k12[P] e t kn + k13 [02]

(16)

where kJ2, kl3, and k11( are rate constants for reactions 12-14, respectively, and [02] is oxygen concentration. The maximum value of φ6ΐ is the ratio of pollutant to oxygen concentration. Because oxygen concentrations are normally constant in the photic zone, sensitized photolysis, like direct photolysis, also obeys a first order rate expression.

Numerous studies have shown that the rate constant for triplet energy transfer (eq. 12) from a sensitizer is determined mainly by the energy of its lowest triplet state, not by its molecular structure (Ref. 18). Energy transfer occurs rapidly only when the triplet energy of the sensitizer equals or exceeds that of the energy acceptor. For example, the rate of photosensitized decomposition of phenylmercury fungicides is rapid with the high energy sensitizer, acetone, but very slow when sensitizers with lower triplet energies are employed (Ref. 29). Listed in Table 2 are triplet state energies of a variety of pesticides, pesticide derivatives, and other chemicals; in most cases the energies were computed from spectroscopic data in the literature.

s 3s*


The excited state energy of molecular oxygen is sufficiently low that triplet sensitizers that absorb uv and visible light transfer energy to it at the maximum rate. The product of energy transfer to oxygen is singlet oxygen, an excited form of oxygen that is a more powerful oxidizing species than ordinary oxygen (Ref. 18). Recently, the wavelength dependence, i.e. action spectrum, for generation of singlet oxygen was found to be strikingly similar in two natural water samples (Ref. 28) (Fig. 4). Because oxygen is the primary energy acceptor in natural waters, the curves in Fig. 4 also approximately represent the variation in efficiency of triplet sensitizer formation with wavelength. In view of what is known about dissolved matter in natural water, it is reasonable to assume that the sensitizers are principally complex, humic substances derived from plant decay (Refs. 30,31). The wavelengths of sunlight absorbed by the sensitizers (Fig. 4) correspond to triplet energies ranging from about 40 to 80 kcal/mole. Because the triplet state energies of many chemicals in Table 2 also fall in this range, photosensitized reaction via energy transfer may be an important transformation process. Energy transfer sensitization is more probable in the case of energy acceptors with lower triplet energies such as oxygen, polycyclic aromatics, parathion, carbaryl, and DDE. The low triplet energies of these compounds indicate that a wide range of sensitizers, including some that absorb visible light, can transfer energy to them.

O.IO

0.08

0.06

0.04

0.02 h

O 0KEFEN0KEE SWAMP ■ AUCILLA RIVER

300 350 400 450 500 WAVELENGTH, NM

550

Fig. 4 Wavelength dependence of the quantum efficiency, Ys, for photosensitized gen-eration of singlet oxygen in samples from two water bodies

Several studi energy transf decomposition correlation b triplet state and coworkers photosensitiz energies of b mole"1) are 1

es in the literature indicate that er are also responsible for photos

Ivie and Casida (24), for examp etween the effectiveness of severa energies. Moreover, Plimmer and (33) have reported that chloroani

ed by benzophenone and riboflavin. enzophenone (68 kcal mole"1) and r ower than those of anilines (> 75

mechanisms other than ensitized le, found a poor 1 sensitizers and their Kearney (32) and Rosen line decomposition is Because the triplet

iboflavin (47 kcal kcal mole"1)(Table 2),


mechanism(s) other than energy transfer must account for these observations.

TABLE 2 Triplet State Energies of Selected Synthetic Chemicals (Refs. 6a,18)

Pesticide or related

DDT Methoxychlor DDE DMDE Parathion Diazinon Carbaryl 2,4-D butoxyethyl ester p-Nitrophenol T-Naphthol Hexachlorobenzene

Triplet Energy

kcal/mole

79 80 54 53 58 76 60 77 58 60 70

Other

Benzene Phenol Aniline Acetone Benzophenone Biphenyl Quinoline Naphthalene Nitrobenzene Chrysene Pyrene Riboflavin Benz (a)anthracene Anthracene Benzo(a)pyrene Naphthacene Oxygen

Triplet Energy

kcal/mole

85 82 11 80 69 65 62 61 60 57 49 47 47 42 42 29 23

Chemical reaction. Another possible mechanism for photosensitization involves chemical reaction between the electronically excited sensitizer and the pollutant (eq. 17).

Products (17)

Of the variety of reactions that can occur, hydrogen atom transfer from pollutant to sensitizer is one of the most general (eq. 18).

S* + P-H > SH' + P· (18)

In the presence of oxygen, this reaction can lead to oxidation of the pollutant by a free radical mechanism. The effectiveness of rotenone as a sensitizer (Ref. 24) is probably related to the fact that it is a substituted phenyl ketone; photoexcited phenyl ketones are known to be efficient hydrogen-atom abstractors (Ref. 34). Riboflavin, a widely used pesticide photosensitizer, is also a good hydrogen abstractor in the excited state (Ref. 35a). The riboflavin-sensitized oxidation of 2,4-dichlorophenol apparently involves hydrogen abstraction (Ref. 35b) as products resulting from coupling of free radicals are obtained.

Excited sensitizers such as ketones can form reactive excited-state complexes, or "exciplexes," with good electron donors (Refs. 36,37) such as alkyl amines (Ref. 38). Oxidation of the amine and reduction of the sensitizer occurs via electron transfer within the exciplex (eq. 19-21).

6" 6 S* + R2NCH2R ) (S··-R2NCH2R)* (19)

Exciplex

S* + P


Exciplex > S" + R2KCH2R > SH· + R2tiCHR (20)

R/NCHR ^ » R2NH + RCOH (21) 02

This mechanism may account for the acetone-sensitized dealkylation of triazine herbicides reported by Burkhard and Guth (39).

Photosensitized oxidation. The combination of light, oxygen, and natural photosensitizers in aquatic environments can lead to photooxidation of a wide variety of substances (Ref. 19). Two general types of sensitized photooxidations can occur (Ref. 40). Type I reactions are initiated by H-atom transfer (eq. 18) or electron transfer (eq. 20) followed by subsequent reaction of the resulting free radicals with oxygen or other chemical species. Type II reactions initially involve the previously mentioned interaction between triplet sensitizer and oxygen, usually to form singlet molecular oxygen (eq. 13). Reactions involving singlet oxygen are often called "oxygenations." The rates and chemistry of both types of reactions have been discussed in an excellent review by Foote (19) , so emphasis here is on our findings concerning singlet oxygen in natural water (Ref. 41).

To obtain quantitative data that can be used to predict rates of photosensitized oxygenations, 2,5-dimethylfuran (DMF) was added to natural water samples to chemically snare the singlet oxygen. Although unreactive in distilled water, the DMF photooxidized in natural waters to form a hydroperoxide that, in turn, decomposed to eis -diacetylethylene and hydrogen peroxide (Fig. 5). That singlet oxygen, and not some other oxidizing species, was involved in the DMF oxidation was established by two mechanistic tests (Ref. 18). The photoreaction was accelerated upon addition of deuterium oxide to the natural water (Ref. 42), and it was slowed when DABCO, a singlet oxygen quencher, was added (Ref. 43). Studies employing monochromatic light (366 nm) established that singlet oxygen is photogenerated in a

l3-C ^>-CH3 CH3<" J>-CH3 XT 0

C i s - Diacetylethylene

Fig. 5 Reaction of singlet oxygen with 2,5-dimethylfuran in water


wide variety of natural waters with a surprising uniformity of quantum efficiency, Ys (Table 3). Steady-state concentrations of singlet oxygen [02*] were estimated from photooxygenation rates of DMF dissolved in natural water exposed to winter sunlight (Ref. 41). The concentrations of 02* generated by sunlight were approximately proportional to the rate of generation of singlet oxygen at 366 nm, as reflected by the product of the fraction of light absorbed, F , and the quantum efficiency, Ys, (Fig. 6). Because Ys is relatively constant in these waters, the variation in [02*] was primarily governed by fraction absorbed. Figure 6 points out an important difference between direct and sensitized photolysis: Sensitized photolysis proceeds most rapidly in water bodies that are most opaque to sunlight7""whereas the opposite is true of direct photolysis. That [02*] generated by exposure to sunTTght correlated with the laboratory data obtained at a single wavelength suggests that the action spectra for photosensitized formation of 02* must be similar in the natural waters that we examined. Indeed, we did find a striking similarity in the action spectra for the Okefenokee Swamp in south Georgia and the Aucilla River in north Florida (Fig. 4).

TABLE 3 Kinetic Parameters for Photochemical Generation of Singlet Oxygen in Natural Waters from the Southeastern United States (Ref. 41)

Water Body

Okefenokee Swamp, GA Fenholloway River, Foley, FL

St. Marks River, St. Marks, FL

Gulf of Mexico, Shell Point, FL

Puddle in peanut fie! Sylvester, GA

Mississippi River, Baton Rouge, LA

Average for 11 water

Ld,

bod

pH

4.1 7.7

8.7

8.5

6.6

8.3

ies

Total Organic Carb< Dn,

24 77

9

6

12

14

mg/1 F a

*366

0.389 0.941

0.165

0.050

0.113

0.060

0.

s 36

0.056 0.014

0.030

0.060

0.065

0.093

043^0.019

(a) Fraction of 366-nm light absorbed in 1.00cm. (b) Quantum efficiency for singlet oxygen formation at 366 nm.

Numerous rate data concerning reactions of singlet oxygen (eq. 22) are available in the chemical literature.

02* + A — £ > A02 (22)

The half-lives for these photosensitized oxygenations equal 0.693 (kA[0 *])

_ 1 (Ref. 41). Literature data concerning k^ values are translated in Table 4 into near-surface half-lives for photosensitized oxygenation of polycyclic aromatics (Ref. 44), biologically important substances (Ref. 45), and some pesticides. Photosensitized oxygenation of pesticides containing oxidizable moieties may also occur rapidly in natural water. A furan derivative, eis-resmethrin

FF

F


(Ref. 46), must photooxidize about as rapidly as DMF. Moreover, Draper and Crosby (47) recently reported that disulfoton, a dialkyl sulfide, is photooxidized to disulfoton sulfoxide in a field water. Oxidation of sulfides is another known singlet oxygen reaction (Ref. 48).

5 ro * O

, Ο

20

16

12

1

1 8

4

o

-

-

-

J L 1 1

o /

1 1 1 1

/ ° f O

1 I L-8 12 16

F 3 6 6 Y S X ' 0 3 20

Fig. 6 Steady-state concentration of singlet oxygen generated by exposure of natural waters to sunlight versus laboratory data obtained using 366 nm; F ? ^ and Ys are fraction absorbed and quantum yield for singlet oxygen generation, respectively.

TABLE 4 Predicted Near-surface Half-lives for Photosen-sitized Oxidation of Selected Organics -- Mid-winter, Southern United States

Polycyclic Aromatics

Benzo(a)pyrene 7,12-Dimethylbenz(a)anthracene Naphthacene Pentacene 2,5-Dimethylfuran

Half-life, hrs

2500 6 7 0.02 0.20

Natural Chemicals

Histidine Tryptophan Methionine Methyl oleate Methyl linoleate

2 2 3

1100 700

Pesticides

cis-Resmethrin Disulfoton

0.2 3


We should note that, although singlet oxygen does rapidly oxidize certain types of chemicals, it is a highly selective oxidizing species. Thus, many of the functional groups widely employed in pesticides are not readily oxygenated (Ref. 26). On the other hand, singlet oxygen may also play an indirect role in photooxidations. Peroxides formed by reactions between singlet oxygen and natural substances (Table 4) may serve as initiators for free radical oxidation (Ref. 6b). Draper and Crosby (47) and Ross and Crosby (25) have reported that peroxides were formed at low concentrations when samples from a variety of water bodies were exposed to sunlight. Ross and Crosby have suggested that these peroxides or "thermal oxidants" are involved in the photosensitized epoxidation of the insecticide, aldrin, to form dieldrin in a natural water (Ref. 49). Singlet oxygen is not directly involved in the reaction. More recently, Shimizu and Bartlett, in a detailed study of the photosensitized oxidation of olefins, suggested that the epoxidations may involve sensitizer-oxygen exciplexes or the biradical species shown below (Ref. 50).

C

R ^ 00·

Biological sensitization. Most of the above considerations concerned photosensitization in natural waters in which non-living humic material was the principal sunlight absorber. O'Kelley and Deason (51) recently discovered an interesting example of biologically mediated photodegradation (Fig. 7). The insecticide, malathion, is not degraded by the action of sunlight when dissolved in pure water (Ref. 6a). However, when aqueous solutions of malathion containing algae (either Chlovella sp. or Nitzsahia sp.) were exposed to visible light, the malathion was rapidly converted to malathion monoacid. Control experiments showed that both light and algae were required for degradation to occur. These observations may be very significant, because algae are the principal sunlight absorbers in many clear lakes and the sea (Ref. 52).

1.0

0.8

[p] ΤΊ °·6 Mo

0.4

0.2

4 8 12 16 20 24

TIME, HRS. Fig. 7 Malathion degradation photosensitized

by algae

Chlorella illuminated


LIGHT ATTENUATION AND PARTITIONING

One of the major uncertainties concerning photolysis rates in the aquatic environment has stemmed from lack of knowledge about the dose of light received by pollutants. In this portion, we discuss techniques for mathematically describing the effects of light attenuation and partitioning upon photolysis rates.

Light Attenuation

Much of what is known about the transmission of solar radiation in natural waters, including ultraviolet light, has been discussed in two excellent reviews (Refs.52,53). In the above discussion, the authors noted that uv and blue solar radiation is mainly responsible for photolysis of man-made chemicals. Most field measurements of underwater light have focused on visible radiation, because it is involved in photosynthesis and is easier to measure. More definitive data on transmission of uv radiation should soon be available, but until it is, only rough estimates based on laboratory data can be made.

Underwater light consists of three parts, direct light from the sun, diffuse light from the sky, and scattered light produced by interaction of light with suspended particles in the water. Light attenuation results from both absorption and scattering, but scattering is known to be less important than absorption in most natural waters. Miller found that suspended freshwater sediments attenuate middle uv radiation (280-340 nm) primarily through absorption (Ref. 16). Thus, as a first approximation, we have ignored scattering in assessing the effects of light attenuation on photolysis (Ref. 9).

Most observers are aware of the diversity in the optical properties of water bodies. Many lakes and the ocean are clear and blue, eutrophic lakes are often green because of suspended algae, and muddy rivers are brown and opaque. The popular belief that water, itself, strongly absorbs uv light, is far from the truth. The photic zone, defined as the depth corresponding to 99% light attenuation, has been shown to exceed 30 meters for uv-B radiation (280-320 nm) in clear ocean water near Puerto Rico and in the Sargasso Sea (Ref. 52). The photic zone for near ultraviolet light (350-400 nm) in the Sargasso Sea and Crater Lake is nearly 100 meters (Refs. 52,54). Many aquatic pollutants, such as polycyclic aromatics, absorb strongly in the near ultraviolet. Light attenuation in coastal or inland waters is usually caused almost completely by dissolved or suspended matter in the water. The photic zone in freshwaters can range from a few centimeters to the above-mentioned maximum values for Crater Lake. The variability in the effects of light attenuation on the direct photolysis of benzo[a]pyrene at latitude 50°N are illustrated in Fig. 8. The average half-lives in the figure were computed assuming absorption coefficients for pure water measured by Dawson and Hulburt (55) and using the average attenuation coefficients for ten river waters in the southern U.S. (Ref. 9). The half-lives were calculated by computer using the program described by Zepp and Cline (Ref. 9).


DEPTH, METERS

Fig. 8 Predicted average half-lives for photolysis of benzo(a)pyrene in water bodies of various depths (midsummer, latitude 50°N)

Effects of Partitioning

Considerations in the preceding section do not take into account the effects of partitioning to sediments. It is well known that certain hydrophobic pollutants such as DDT and benzo[a]pyrene have a great tendency to sorb to both suspended and bottom sediments. Because most of the sorbed pollutant is in the bottom sediment where no light is present, water-sediment partitioning decreases photolysis rates in aquatic environments. The tendency to sorb can be quantitatively described by the partition coefficient, Kp. The value of Kp is numerically equal to the weight ratio of sorbed to dissolvea pollutant in a system that contains equal amounts of water and sediment. Figure 9 shows that the primary effect of suspended sediments on photolysis is the result of light attenuation, not partitioning. Typically, suspended sediment concentrations are 50 mg/liter or less (Ref. 16), and most pollutant in the water column is in the dissolved, not sorbed, form. Moreover, available evidence (Refs. 15,16) suggests that photolysis rates of sorbed and dissolved pollutant are similar.

To assess the effects of sorption to bottom sediment, equations shown in Fig. 10 were derived. The terms [Ps] and [Pw] represent concentrations of pollutants in sediments and water column, respectively, and Cs is the ratio of sediment to water. In deriving the equation, we assume that sediment-water exchange is much more rapid than transformation. If the pollutant is transformed much more slowly in the sediment than in the water column, i.e., if ks << kw, then the observed photolysis rate is reduced by a factor of (1 + KpCs)

_1 relative to the rate in absence of partitioning. The value (1 + KpCs)

_1 incidentally, also corresponds to the fraction of pollutant in the water column. The computed effect of partitioning upon loss of DDE by photolysis and volatilization represents an extreme example


(Fig. 11). The volatilization rate of DDE was computed as described by Mackay and Leinonen (Ref. 56). It is reasonable to assume that partitioning changes the rate of breakdown of pollutants by other pathways as well. For example, Juengst and Alexander have suggested that the microbial breakdown of DDT is slowed by partitioning (Ref. 57). Thus, chemicals with large partition coefficients may have a great tendency to persist, as well as bioaccumulate (Ref. 58), in aquatic environments.

% of UV Attenuated in 10 cm. (KAtten =0.0026 l/mg-cm)

Attend C t

-Lo ' S

% of Chemical Sorbed

20 40 60 80 100 120 140 160

SEDIMENT CONCENTRATION (mg/l)

Fig. 9 UV light attenuation and sorption by suspended sediments, attenuation concentration (mg/l), respectively; Kp is the sediment-water partition coefficient. Value of ΚΑ££θη is average for sediments obtained from five water bodies in Georgia, U.S.A. (Ref. 17)

The terms, KAt1-en, d, and C~ are coefficients, depth, and sediment

Products or Physical Loss

w W- * Ps —Ξ-— Products

[Psl p CS[PW]

Cs = Sediment Concentration,g/l

[ P T ] = [ p w ] + [ ps ]

d[PT] kw + k sC s K P

d t l+KpCs [PT]

W

I + K P C S

[PT] if k s is

very low.

Fig. 10 Equations describing effects of parti-tioning upon transformation


100

10

■o

< x

0.1

SEDIMENT PRESENT

CONDITIONS ASSUMED: •20 MG/L SUSPENDED SEDIMENT •ACTIVE LAYER OF BOTTOM SEDIMENTS

1.5 CM DEEP • PARTITION COEFFICIENT = 7000

NO SEDIMENT PRESENT

J ' I I I l I l l J 1 I I I I I I 1 0

DEPTH, m 100

Fig. 11 Computed effects of partitioning upon photolysis and evaporative loss of DDE from water bodies of various depth

The possibility has pollutants in surfa formation (Ref. 59) the Great Lakes (Re trace metals and ch fold greater in the in surface films se concentration of po amount of pollutant even in just 1 mete reached by Singmast

also been suggested that accumulation of ce films can accelerate photochemical trans-

Analysis of slicks from the sea (Ref. 60) and f. 61) have confirmed that the concentrations of lorinated chemicals such as PCBfs are 10- to 20-top 100 ym than in underlying water. Photolysis ems unimportant to us, however. Although the llutant is higher in the surface film, the total in the film is negligible compared to the amount

r of underlying water. This conclusion was also er after more detailed study (Ref. 62).


OUTDOOR VERIFICATION

In conclusion, we present some experimental evidence that our attempts to predict photochemical transformation are on the right track. During the past few years, we (Refs. 6a,9) and Smith and coworkers (5) have conducted numerous experiments in Georgia and California designed to compare predicted direct photolysis half-lives with half-lives measured under sunlight. The experiments were mainly conducted at shallow depths in pure water with no sediment present, although no appreciable changes in half-life were detected when natural water from oligotrophic water bodies was used. The computed and experimental half-lives were found to be in close agreement for pesticides and polycyclic aromatic compounds (Fig. 12). Presently, we are developing methods for predicting effects of sediments and natural photosensitizers upon photolysis rates.

3 lh < IE Q

O I

< O O o

oh

1 OZepp and Cline.1976 | D M a b e y and Mill, 1976

7o \- 6 o '

,05 D l 5

W P. , 1

/

1 - 1 0 1 2

LOG OBSERVED HALF-LIFE, hrs

Fig. 12 Comparison of computed half-lives for direct photolysis of chemicals in water and half-lives measured at shallow depths under sunlight. The numbers correspond to chemicals listed in Table 1.

Usually, the behavior of a pollutant in a field situation is the result of a number of physical, chemical, and biological interactions. Modeling the behavior of a pollutant, however, is simplified when one or two processes are exceptionally rapid. For example, predicting the loss of vinyl chloride (Ref. 4) or chloroform (Ref. 63) from aquatic systems is greatly simplified by the dominance of volatilization. Laboratory data suggest that the major processes responsible for loss of DDE from the aquatic environment are photolysis and volatilization (Refs. 8,64) (Fig. 11). Experiments have shown that DDE is biodegraded very slowly (Refs. 1,65) and is chemically stable (Ref. 66). During June 1972, Hamelink and Waybrant (65) started a field study of the behavior of DDE in a large quarry in southern Indiana. Of particular interest to us was the unexplained loss of DDE from the quarry during the first year (Fig. 13). The initial rapid loss from the water column during the first 20 days was in part accounted for by


sorption to the bottom sediment. After this initial period, loss occurred much more slowly. Figure 13 shows the remarkable agreement between the observed concentration-time profile for DDE and the profile computed from laboratory data (Fig. 11) for a 15-meter-deep oligotrophic water body.

O OBSERVED (HAMELINK AND WAYBRANT. 1973) D COMPUTED FOR 15.5-M DEPTH

, I i I I I I I I 0 20 40 60 80 100 120 140

DAYS

Fig. 13 Loss of DDE from the water column of an oligotrophic quarry at Oolitic, Indiana, during Summer 1972.

CONCLUSION

A technique is now available for using laboratory data to predict rates of direct photochemical reactions of aquatic pollutants. The technique allows estimates of the variations in photolysis rates as a function of location, time, and light attenuation in water bodies. Direct photolysis is likely to be an important pathway for trans-formation of chemicals such as pesticides, polycyclic aromatic hydrocarbons, and nitrosamines. Photolysis rates are expected to vary widely from one water body to another due to the effects of light attenuation and partitioning. The decrease in direct photolysis rates with increasing turbidity may be offset by an increase in rates of photosensitized reactions. Absorption of sunlight by natural photosensitizers can lead to transformation of chemicals that are photochemically inert in pure water. Thus, direct photolysis rates represent minimum values for rates of photochemical transformation in the aquatic environment.

ACKNOWLEDGEMENTS

The technical assistance of P. Schlotzhauer and R. Hollis is grate-fully acknowledged. D. M. Cline assisted in writing the computer program that was used to calculate direct photolysis rates. Some of the data used in computing photolysis rates were obtained by W. R. Mabey, T. Mill, and their coworkers at SRI International, Menlo Park, CA.


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N

C

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D

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(36) R. M. Bowman, T. R. Chamberlain, C. Huang, and J. J. McCullough, Medium effects and quantum yields in the photoaddition of naphthalene and acetonitrile. Chemical evidence on an exciplex structure, J_ Amer. Chem. Soc. 96, 692 (1974).

(37) P. J. Wagner and A. Kemppainen, Solvent effect on a probable charge-transfer reaction. Inter and intramolecular photo-reactions of tertiary amines with ketones, J. Amer. Chem. Soc. 91, 3085 (1969).

(38) S. G. Cohen, Photoreduction by amines, Chem. Rev. 73, 141 (1973).



(39) N. Burkhard and J. A. Guth, Photodegradation of atrazine, atraton, and ametryne in aqueous solution with acetone as a photosensitizer, Pestic. Sei. 7, 65 (1976).

(40) K. Gollnick, Type II photooxygenation reactions in solution, Advan. Photochem. 6, 1 (1968).

(41) R. G. Zepp, N. L. Wolfe, G. L. Baughman, and R. C. Hollis, Singlet oxygen in natural waters, Nature 267, 421 (1977).

(42) P. B. Merkel and D. R. Kearns, Radiationless decay of singlet molecular oxygen in solution, J. Amer. Chem. Soc. 94, 7244 (1972).

(43) C. Ouannes and T. Wilson, Quenching of singlet oxygen by terti-ary aliphatic amines. Effect of DABCO, J. Amer. Chem. Soc. 90, 6527 (1968).

(44) B. Stevens, S. R. Perez, and J. A. Ors, Photoperoxidation of unsaturated organic molecules, J. Amer. Chem Soc. 96, 6846 (1974).

(45) R. Nilsson, P. B. Merkel, and D. R. Kearns, Unambiguous evidence for participation of singlet oxygen in photodynamic oxidation of amino acids, Photochem. Photobiol. 16, 117 (1972).

(46) K. Ueda, L. C. Gaughan, and J. E. Casida, Photodecomposition of resmethrin and related pyrethroids, J^ Agr. Food Chem. 22, 212 (1974).

(47) W. M. Draper and D. G. Crosby, Measurement of photochemical oxidants in agricultural field water, presented in part at the 172nd National Meeting of the American Chemical Society, San Francisco, September 1976.

(48) C. S. Foote and J. W. Peters, Chemistry of singlet oxygen. A reactive intermediate in sulfide photoxidation, J^ Amer. Chem. Soc. 93, 3795 (1971).

(49) R. D. Ross and D. G. Crosby, The photooxidation of aldrin in water, Chemosphere 4, 227 (1975).

(50) N. Shimizu and P. D. Bartlett, Photooxidation of olefins sen-sitized by α-diketones and by benzophenone, J^ Amer. Chem. Soc. 98, 4193 (1976) .

(51) J. C. O'Kelley and T. R. Deason, Degradation of Pesticides by Algae, U.S. Environmental Protection Agency Report EPA-600/3-76-022, Athens, GA, March 1976.

(52) R. C. Smith and J. E. Tyler, Transmission of solar radiation into natural waters, Photochem. Photobiol. Rev. 1, 117 (1976).

(53)J. R. V. Zaneveld, Penetration of ultraviolet radiation into natural waters, in Impacts of Climatic Change on the Biosphere, CIAP Monograph 5XT975) , U.S. Department of


Transportation, Climatic Impact Assessment Program, Report DOT-TST-75-55, Washington, DC, p. 2-108.

(54) R. C. Smith, J. E. Tyler, and C. .R. Goldman, Optical properties and color of Lake Tahoe and Crater Lake, Limnol. Qceanog. 18, 189 (1973).

(55) L. H. Dawson and E. 0. Hulburt, The absorption of ultraviolet and visible light by water, J^ Opt. Soc. Amer. 24, 175 (1934).

(56) D. Mackay and P. J. Leinonen, Rates of evaporation of low-solu-bility contaminants from water bodies to atmosphere, Environ. Sei. Technol. 9, 1178 (1975).

(57) F. W. Juengst and M. Alexander, Conversion of 1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane (DDT) to water-soluble products oy microorganisms, J. Agr. Food Chem. 24, 111 (1976). —

(58) C. T. Chiou, V. H. Freed, D. W. Schmedding, and R. L. Kohnert, Partition coefficient and bioaccumulation of selected organic chemicals, Environ. Sei. Technol. 11, 475 (1977).

(59) R. Rabson and J. R. Plimmer, Science 180, 1204 (1973).

(60) R. A. Duce, J. G. Quinn, C. E. Olney, S. F. Piotrowicz, B. J. Ray, T. L. Wade, Enrichment of heavy metals and organic compounds in the surface microlayer of Narragansett Bay, Rhode Island, Science 176, 161 (1972).

(61) A. W. Elzerman and D. E. Armstrong, Surface organic microlayer-microcontaminant interactions in Southern Lake Michigan, 39th Annual Meeting of the American Society of Limnology and Oceanography, Savannah, GA, June 1976.

(62) J. Singmaster, Environmental behavior of hydrophobic pollutants in aqueous solutions, Ph.D. dissertation, University of California, Davis, CA, 1975.

(63) W. B. Neely, G. E. Blau, and T. Alfrey, Jr., Mathematical models predict concentration-time profiles resulting from chemical spill in a river, Environ. Sei. Technol. 10, 72 (1976).

(64) R. G. Zepp, N. L. Wolfe, L. V. Azarraga, R. H. Cox, and C. W. Pape, Photochemical transformation of the DDT and methoxy-chlor degradation products, DDE and DMDE, by sunlight, 172nd National Meeting of the American Chemical Society, San Francisco, September 1976; Arch. Environ. Contam. Toxicol. in press (1977).

(65) R. L. Metcalf, J. R. Sanborn, L. Po-yung, and D. Nye, Laboratory model ecosystem studies of the degradation and fate of radiolabeled tri-, tetra-, and pentachlorobiphenyl compared with DDE, Arch. Environ. Contam. Toxicol. 3, 151 (1975).

(66) N. L. Wolfe, R. G. Zepp, D. F. Paris, G. L. Baughman, and R. C. Hollis, Methoxychlor and DDT degradation in water: rates and products, Environ. Sei. Technol. in press (1977).


(67) J. L. Hamelink and R. C. Waybrant, DDE and Lindane in a large-scale model lentic ecosystem, Trans. Am. Fish. Soc. 124 (1976).

Behaviour of Persistent Organic Compounds in Bank-filtrated Rhine Water

HORST KUSSMAUL

Institute for Water, Seil and Air Hygiene, Federal Health Office, 6000 Frankfurt/M., Federal Republic of Germany

ABSTRACT

Almost one half of the drinking water derived from surface water in the Federal Republic of Germany is obtained from ground water replenished by way of bank fil-tration. Where water works using bank filtrate and periods of several month for storage of the infiltrated water in the subsoil, ground water quality is reached; in the case of waiting periods of only a few days or weeks, however, part of the organic matter is eliminated only insufficiently. Above all, a description is given of results obtained in a water works at the lower Rhine using bank filtrate with an average storage time of 20 days and a median proportion of river water within the raw water obtained of 80 %.

In the immediate vicinity of the bank , dissolved organic matter is decomposed up to 60 % while the compounds which are difficult to decompose will hardly undergo any further alteration during onward subsoil passage. For a more detailed charac-terization of organic matter, group parameters, such as volatile and non volatile extractable organochlorine and organofluorine compounds, cholinesterase inhibitors and aromatic amines have been determined. It was revealed that in the course of bank filtration, these compounds were retained only by an average of 60 to TO %. However, defined individual compounds from the above-mentioned groups may behave quite differently as is shown by a number of examples.

Subsequent treatment of raw water for drinking water production by means of ozoni-zation and three-stage filtration over activated carbon is capable of eliminating a major proportion of difficult to decompose organic substances from the water.

INTRODUCTION

Because of its balanced characteristics, ground water is preferred as a source of drinking water supply. At least in the Federal Republic of Germany, the utilizable ground water resources are being exploited to a considerable degree. In the last decades, these resources have not been able any more to satisfy the rapidly growing water requirements as a consequence of progressing industrialization and increased private use with the onward refinement of civilization, so that increa-sing recourse had to be made to surface waters as a source for the production of drinking water. This development will continue into the future even though the rates of water consumption increase will not be as high as they were assumed to become some years ago.

265

266 H. Kussmaul

Within the total area of the Rhine basin, 20 million inhabitants are to be supp-lied with drinking water; the raw water needed for 8 5 million of them is taken from the Rhine. In 1973, approximately 620 million m of water were taken from the Rhine for purposes of drinking water supply (the amount of water taken for use as industrial and cooling water was exceeding ten times this amount), 12 % thereof by Switzerland, 39 % by the Federal Republic of Germany, and h9 % by the Nether-lands (1).

In contrast to a number of other countries, there is hardly any immediate treat-ment of river water in the Federal Republic of Germany; rather, this water is used for ground water replenishment. By this procedure, a pretreatment of the river water is achieved by taking advantage of natural biological, chemical, and physi-cal processes taking place in the geological subsoil.

Ground water replenishment may take place by way of recharge basins and injection wells or bank filtration. Natural bank filtration will be present where there is a hydraulic gradient between the level of the surface water and the ground water. It may be intensified by the formation of a pumping depression cone near the bank as a consequense of water intake (2).

Almost one half of the drinking water which is derived from surface water sources in the Federal Republic of Germany is obtained by bank filtration for ground water replenishment (3). Most of the water works using this method are situated in the lowlands of the large rivers, above all of the Rhine.

Where the infiltrated water is left in the subsoil for some month before it is taken for treatment in those water works at the lower Rhine which utilize bank filtration, ground water quality is achieved while storage periods of a few days or weeks only will result in an insufficient elimination of parts of the organic substances contained will be shown by a number of examples.

DESCRIPTION OF WATER WORKS

Above all, an account will be given of the results obtained in a water works situated at the lower Rhine and using bank filtrate. The battery of wells is si-tuated parallel to the river bank at about kO m distance from the outer (flow) side of a loop of the Rhine. 78 vertical wells are used for the intake of the water.

For the purpose of a hydrochemical determination of the periods during which the bank filtrate remained in the subsoil between infiltration and intake, the amounts of chloride and borate ions which are present in the Rhine with variations in the course of time proved to be well suitable. Studies in these water works over a number of years revealed average monthly storage periods of 20 days with varia-tions between 3 and k6 days. Regression analysis has shown that there is a close linear correlation between the reciprocal value of the storage period and the level of the Rhine (U).

Sodium and chloride ions are well suitable for a determination of the proportion of river water in the raw water taken in. In the case of the water works described, the average proportion of river water is one of 80 % with individual values vary-ing between 30 and 100 %. Also the proportion of river water has been exhibiting a close correlation to the river level (h),

Because of the brief storage times of the bank filtrate in the subsoil, an expen-sive system of water treatment including ozonization and a three-stage filtration over activated carbon has been additionally installed for the purpose of onward treatment of this raw water.

Behaviour of Persistent Organic Compounds 267

CHARACTERIZATION OF ORGANIC MATTER

For the general characterization of the organic substance contained in the Rhine water, the dissolved organic carbon (DOC), the chemical oxygen demand (COD), and the potassium permanganate consumption, as cumulative parameters, were determined. The KMnO, consumption will provide a certain indication of the amount of easily decomposable organic matter while COD also takes into account substances which are more difficult to decompose.

Our studies over a number of years revealed that for this water works using bank filtrate from the Rhine the underground passage (taking into account the propor-tion of ground water and its concentrations of organic matter) reduced the DOC of the river water only by an average of 56 %9 the COD by 66 %9 and the KMnO, con-sumption by 71 % (Fig. 1). By detailed studies it was found that elimination of the organic substances is taking place primarily in the immediate vicinity of the river bank (5).

KMnO^-Verbr.

Rh Uf Tw Gw Rh Uf Tw Gw Rh Uf Tw Gw

Fig. 1. Dissolved organic carbon, chemical oxygen demand, and potassium perman-ganate consumption in the Rhine (Rh), bank filtrate (Uf), drinking water (Tw), and ground water (Gw), mean values from April 1972 up to January 1975

Intensive treatment of the water for the production of drinking water is only capable of reducing the DOC and the KMnO, consumption by another 10 % and the COD by another 20 % (Fig. 1). It is seen from the DOC contents that on the average

268 H. Kussmaul

one third of the dissolved organic matter in water from the Rhine is neither eli-minated during underground passage nor in the process of drinking water production. However, to ensure a safe drinking water supply, an onward study of these persist-ent substances is required. For this purpose, group parameters were determined and also individual substances identified.

ORGANOCHLORINE COMPOUNDS

Organochlorine compounds do not naturally occur in the environment. However, they are produced and used in large quantities, thus reaching surface waters through effluents. Several of these substances are known for their persistence and capa-bility of accumulation in the environment. It is thus important to follow up, above all, the behaviour of organochlorine compounds in the process of bank fil-tration.

The analysis for organochlorine compounds is performed either by microcoulometry or by means of an ion-specific electrode with preceding pyrolysis. Analysis covers easily volatilized and not easily volatilized extractable compounds. The volatile compounds are identified by degassing and direct introduction into the pyrolysis oven, the non volatile ones following extraction with diisopropyl ether (6).

TABLE 1 Behaviour of organochlorine compounds during bank filtration

Organochlorine compounds (yg/l)

volatile non volatile

Rhine 1+0.6 13.8

Bank filtrate 19.9 h.2

Drinking water 8.6 0.9

In Table 1 the mean values from 6 test series for the extractable non volatile substances and 3 test series for the volatile substances are given. It is seen that in the course of bank filtration only about one half of the volatile sub-stances is retained. Following treatment by ozone and multi-stage filtration over activated carbon still more than 20 % as compared with the concentration in river water are present. The elimination of the non volatile extractable substances is somewhat better: during bank filtration, they are eliminated from the water by 70 % and during onward treatment by more than 90 % (6).

At this moment, we are trying to identify the most important individual compounds by means of capillary gas chromatography and mass spectrometry. During this work, it has been found that the foremost amount of the identifiable organochlorine com-pounds in the bank filtrate and the drinking water is formed by the four compounds chloroform, carbon tetrachloride, trichloro ethylene, and tetrachloro ethylene which are present in amounts in the lower microgram range per liter of water. In the process of bank filtration, these compounds are not retained and in the method of onward treatment used only insufficiently.

In contrast to this, numerous individual compounds from the group of the non vola-


tile extractable compounds may be identified in the river water and bank filtrate within the range from a few to several hundred nanograms per liter. As the most important components, primarily dichloro and trichloro benzenes were found in amounts between 300 and 500 ng/l. Chlorinated pesticides and related compounds which are also included in this group are only of minor importance from the quan-titative point of view, provided they are detectable anyway. The behaviour of the individual compounds within this fraction during bank filtration and the treatment of raw water to obtain drinking water is varying considerably.

Since it is impossible to sample river water and bank filtrate with appropriate correspondence in terms of time, and, on the other hand, effluents are discharged intermittently into receiving waters, such studies have to be performed over ex-tended periods to be capable of stating reliable elimination rates for the indi-vidual compounds.

In conclusion, it should be added that the compounds discussed account for only about 10 to 20 ? of total organic chlorine in the river Rhine in accordance with studies by Kühn (7) while very little has become known about the remaining pollu-tion load consisting of polar compounds that cannot be extracted.

ORGANOFLUORINE COMPOUNDS

Since in addition to organochlorine compounds, also considerable amounts of organo-fluorine compounds are produced, their presence in the Rhine and their behaviour during bank filtration will also be considered. Determination is performed by means of a fluorine specific electrode or photometric methods with preceding pyro-lysis, separately for the easily volatilized and the not easily volatilized ex-tractable compounds (8).

The pollution load of the river Rhine due to organofluorine compounds has been depicted in Fig. 2. Evaluation covered five tours of the Rhine from Lake Con-stance to the Dutch border in 1976 and 1977. It is seen that the highest values of about 5 ug/l organic fluorine for both the volatile and non volatile compounds were measured down-stream from the mouth of the Main river while later on values decreased again as a consequence of degassing and dilution. It may be seen from the values measured at the lower Main which have also been plotted, that the main pollution load of the Main and the Rhine rivers due to organofluorine compounds come from one main source, a large industrial plant (9).

For the water works at the lower Rhine using bank filtrate, so far two series of measurements of organofluorine compounds have been available. As a result, vola-tile organofluorine compounds could not be identified at the times of sampling, neither from the river nor from the bank filtrate. Non volatile organofluorine compounds which were found in the river in amounts between 0.1 and O.k yg F/l did not decrease during bank filtration; however, their content in drinking water was below the detection limit.

Of far more concern for drinking water supply is the pollution load of the lower Main due to organofluorine compounds. A water works situated at the middle of the Rhine down-stream from the mouth of the Main at the right bank makes ground water replenishment by water from the central current of the Rhine which does not con-tain organofluorine compounds, which is true also for water which is infiltrated through basins and wells. But the influence of the bank filtrate from the water of the Main river which is passing the bank almost in its original unmixed condi-tion could be clearly felt. Since in this case the organofluorine compounds could be demonstrated to penetrate into the drinking water which has also been the case with a water works on a Rhine island using bank filtrate somewhat more upstream, an identification of the individual compounds is attempted at present, to enable

270 H. Kussmaul

a hygienic eva lua t ion of t h e i r presence in t he dr inking water (9 ) .

ORGANOFLUORINE COMPOUNDS f-3 /jgF/l .2 LI LO ■ NON VOLATILE

□ VOLATILE

Fig. 2. Organofluorine compounds in the river Rhine (and some tributaries), mean values from March 1976 up to January 1977

CHOLINESTERASE INHIBITORS

Because of the considerable limitations for the application of chlorinated insec-ticides, phosphoric esters are increasingly applied for this purpose. The toxic effect of these compounds and, to a lesser degree, of the carbamate insecticides consists in a specific inhibition of the enzyme cholinesterase. Although large concentrations of other substances, such as chlorophenols will inhibit this en-zyme, the cholinesterase inhibition test appears to be suitable as a parameter of effects for evaluation of water quality because of its sensible reaction to phos-phorous pesticides. To determine the current inhibitory effect, the parathion equivalent value without oxidation, and for a determination of the possible inhi-bitory potential, the paraoxon equivalent value with preceding oxidation, were measured (10).

Average parathion equivalent values for the water works at the lower Rhine are plotted in Fig. 3. The calculated elimination rate for the underground passage is one of 72 %. The cholinesterase inhibitors are retained almost exclusively in the immediate bank area.


Parathionäquiv. Paroxonäquiv-

IC?

j/ig«

H?

i

id

i

«P-

jjgA

P:

i

10

In

Rh Uf Tw Gw Rh Uf Tw Gw

Fig. 3. Cholinesterase inhibitors in the Rhine (Rh), bank filtrate (Uf), drinking water (Tw), and ground water (Gw), mean values from September 1973 up to May 1975

Strikingly, the parathion equivalents rose to 200 to 400 yg/l after ozonization in the process of water treatment, and average values around 60 yg/l were not reached before the stage of activated carbon filtration. These values are twice as high as the corresponding ones for ground water.

In the case of the paraoxon equivalent values, an elimination rate of 9^ % for un-derground passage was calculated. After water treatment, no difference to ground water may be recognized by means of this parameter.

Only dimethoate could be regularly demonstrated by gasChromatographie analysis of water samples for the most common pesticides. The substance was present in river water in concentrations between 1 and 10 yg/l, in raw water in concentrations be-tween 3 and 5 yg/l. Also in this case, the retention potential of the subsoil is quite low; the concentration is reduced to one tenth only by water treatment. In addition to dimethoate, parathion and malathion have been occasionally found. How-

272 H. Kussmaul

ever, the amount of the active substances identified could account at best for one percent of the cholinesterase inhibition effect measured, if converted according-ly. The remaining part appears to have been caused by reactive metabolites of the comparatively instable phosphoric esters or other cholinesterase inhibitors un-known so far.

AROMATIC AMINES

As another indicator of the presence of sewage and industrial wastes, the aromatic amines are determined. They may have their origin in wastes from the chemical in-dustries or be products of the decomposition of phenylurea herbicides. Determina-tion is performed by photometry; the results are expressed as 3.^-dichloro aniline equivalents (11).

AronrtAmine Trifluormethylanilin Anilin 4-Chloranilin 3,A-DichloranUin

Rh Uf Tw Rh Uf Tw Rh Uf Tw Rh' Uf Tw Rh Uf Tw

Fig. k. Aromatic amines in the Rhine (Rh), bank filtrate (Uf), and drinking water (Tw), mean values from September 1973 up to September 197^

The measuring results for the water works at the lower Rhine have been plotted in Fig. k. In the immediate vicinity of the bank, the dichloro aniline equivalents are decreasing by about 60 %9 and during the entire subsoil passage by 70 %· By this method, aromatic amines are not detectable in drinking and ground water.

From the number of individual compounds, aniline, 3.^-dichloro aniline, 3-tri-fluoro methyl aniline, and ^-chloro aniline could be regularly identified in the water samples and determined by gas chromatography which revealed decreasing amounts in the order given (Fig. h).


These compounds exhibit a quite varying behaviour during bank filtration. When the mixture ratio in the raw water drawn off is taken in account, there have been cal-culated decreases of 57 % for dichloro aniline, 20 % for aniline, 6 % for chloro aniline, and 1 % only for trifluoro methyl aniline (11).

It is seen from the elimination rates that also these compounds are only insuffi-ciently retained in the course of underground passage. However, the method of treatment applied is capable of retaining almost completely those substances which may still be present in the raw water.

CONCLUSIONS

The production of drinking water by way of bank filtration offers the great advan-tage of low area requirements (which may be of decisive importance for areas of urban agglomerations with high water requirements) and the missing need for in-vestment costs to cover replenishment installations. It has the disadvantage of not beitLg in a position to discontinue with the infiltration of river water if a high pollution load is present in that river so that intermittently, large amounts of pollutants may penetrate into the subsoil from which they cannot be eliminated any more at justifiable financial expenditure. For this reason, instal-lations for the conversion of this raw water into drinking water should be de-signed with an especially high safety margin. For rapid monitoring of the water quality and recognition of any alteration of trends, cumulative and group para-meters are suitable in the case of organic matter. However, for a hygienic and toxicologic evaluation, of the water, it will be necessary to identify the indivi-dual persistent compounds since even substances which are quite similar in their chemical nature may behave quite differently during bank filtration and water treatment.

It is an advantage of bank filtration over direct treatment of river water that the biochemical pretreatment in the subsoil makes breakpoint-chlorination dispen-sable. Although organohalogenide compounds will form also in the process of trea-ting bank filtrates, their amounts will be less than those occuring after break-point-chlorination by more than one tenth power (12, 13).

REFERENCES

(1) H. Hammel, Das Langzeitprogramm der Rheinsanierung und die Trinkwasserversor-gung am Rhein. Ber. 5. Arbeitstagung der Int. Arbeitsgem. der Wasser-werke im Rheineinzugsgebiet, 93-102 (1975).

(2) Fachausschuß "Wasserversorgung und Uferfiltrat" des Bundesmin. des Innern (1975) Uferfiltration, Bonn.

(3) Bundesverband der deutschen Gas- und Wasserwirtschaft (1975) Intern. Wasser-statistik 1968-1972, Frankfurt/M.

(U) H. Kußmaul, D. Mühlhausen und H. Behrens, Hydrol. und hydrochem. Untersuchun-gen zur Uferfiltration, Teil I, gwf - wasser/abwasser 118 (1977) in print.

(5) H. Kußmaul und D. Mühlhausen, Hydrol. und hydrochem. Untersuchungen zur Ufer-filtration, Teil III, gwf - wasser/abwasser, in preparation.

(6) U. Fritschi, G. Fritschi und H. Kußmaul, Mikrocoulometrische Summenbestimmung von schwer- und leichtflüchtigen Organochlorverbindungen im Wasser, in preparation.

H. Kussmaul

W. Kühn, Veränderungen des Gehalts von Organohalogen-Verbindungen im Wasser bei der Trinkwassergewinnung und -aufbereitung, in print.

H. Kußmaul und M. Hegazi, Die Bestimmung von Organofluorverbindungen im Was-ser, Vom Wasser U8, 1U3-151* (1977).

M. Hegazi und H. Kußmaul, Herkunft von Organofluorverbindungen im Oberflächen-wasser und ihre Analytik, in preparation.

G. Fritschi, H. Kußmaul und J. Sonnenburg, Cholinesterase-Hemmtest zur Beur-teilung der Wassergewinnung durch Uferfiltration, Vom Wasser k59 75-90 (1975).

M. Hegazi, Analytik und Verhalten von Phenylharnstoff-Herbiziden und deren Metaboliten bei Uferfiltration, Trinkwasseraufbereitung und Bodenpassa-ge, Diss. Bonn 1977.

L. Stieglitz, W. Roth und W. Kühn, Das Verhalten von Organohalogenverbindun-gen bei der Trinkwasseraufbereitung, Vom Wasser 1*7, 3^7-377 (1976).

D. Maier und H. Mäckle, Wirkung von Chlor auf natürliche und ozonte organi-sche Wasserinhalts Stoffe, Vom Wasser 1+7, 535-397 (1976).

274

( 7 ) V

(Ö) ί

( 9 ) IV

( 1 0 ) G

( 1 1 ) l·

( 1 2 ) I

( 1 3 ) I

Behaviour ofDegradable Chemicals in the River Trent

J. H. N. GARLAND

Water Research Centre, Stevenage, Herts, U.K.

INTRODUCTION

In the United Kingdom much legal significance is attached to the concept of wholesomeness as the attribute of water to be taken from surface and underground sources for public supply. The interpretation of this principle is largely a matter of experienced judgement, and this was exercised in 1972 when it was found that although water from the River Trent could be purified to potable standards it could not be regarded as being wholesome.

Comparison of the quality of Trent water which existed in 1970 with the standards subsequently promulgated by the E.E.C. (1) would suggest that this judgement was well founded. Water in the Trent at that time possessed a quality which was outside the limits for Category A3 specified in the directive issued by the E.E.C. in 1975. Water in this category requires intensive physical and chemical treat-ment, as was found by Miller and Short (2). Doubts about the wholesomeness of the water arose when it was found that despite extensive and prolonged treatment with activated carbon the finished water contained residues of organic carbon whose composition and origin were unknown. As a result it was concluded by Rowntree (3) in 1972 that »the use of the Trent for potable supplies could not be recommended for at least the next decade*.

Because of the interest in the organic carbon content of Trent waters this paper re-examines some of the data collected during the Trent Research Programme (4) with a view to deducing the stability of this group of substances in river waters.

ORGANIC CARBON IN THE TRENT SYSTEM

The annual mean daily mass flows and annual flow-weighted mean concentrations of soluble (SOC) and particulate organic carbon (total organic carbon minus SOC) during 1970 in the principal tributaries in order of joining the Trent above Nottingham are summarized in Table 1.

275

276 J. H. N. Garland

TABLE 1 Mass Flows and Concentrations of Soluble and Particulate Organic Carbon in Some Tributaries of the Trent System 1970

Tributary

Upper Trent

R. Tame

R. Dove

R. Derwent

R. Soar

R. Erewash

Total

Lower Trent

Soluble organic carbon

Mass flow (t/d)

5.9

24.5

5.4

11.1

6·. 8

2.5

56.2

44.3

Concentration (mg/1)

5.2

13.1

4.2

6.6

7.7

17.7

6.5

Particulate organic carbon

1 Mass flow (t/d)

3.0

; 9.1

2.1

4.7

2.3

0.7

21.9

20.0


2.7

4.9

2.4

2.8

2.6

5.0

2.9

Of the rivers listed in Table 1 the Tame and the Erewash are grossly polluted fishless streams. The Rivers Dove and Derwent are abstracted for public supply; 90 000 m /d from the Dove being supplied to the City of Leicester and 26 000 m^/d from the Derwent to the City of Nottingham. Water is taken from the Derwent below a point where 96 000 m^/d of treated domestic and industrial effluent from synthetic fibre and organic chemical manufacturers enter the river, and it has been found necessary to use activated carbon to control the taste and odour of the water put into supply (5). It is interesting to note that the concentrations of organic carbon in the Lower Trent and the R. Derwent were very similar in 1970.

SOURCES OF SOLUBLE ORGANIC CARBON AND RELATIONSHIP TO CARBONACEOUS BOD

The data given in Table 1 show that the important sources of SOC are the Rivers Tame and Derwent which together account for 63.3 per cent of the total input from the tributaries. Since about one-quarter of the water flow in the Lower Trent comes from effluent discharges it is necessary to make some estimate of the contributions from these. An examination of the SOC content of 7 treated effluents discharged to the R. Tame indicated a mean concentration of 22.3 mg/1 (4). It is assumed that such a concentration is typical of all effluents discharged in the basin, and from a knowledge of the mean effluent flow in each tributary the mass flows derived from effluents can be calculated. These values are given in Table 2.

Behaviour of Degradable Chemicals 277

TABLE 2 Observed Mass Flows of SOC and Carbonaceous BOD, Calculated Mass Flows of SOC Derived from Effluent and Other Sources, and

SOC Degradable and Stable in the 5-day BOD Test

Tributary

Upper Trent

R. Tame

R. Dove

R. Derwent

R. Soar

R. Erewash

Total

Lower Trent

Mass flow of

Observed SOC (t/d)

5.9

24.5

5.4

11.1

6.8

2.5

56.2

44.3

SOC from effluents

(t/d)

3.6

18.9

1.4

4.3

4.5

1.4

34.1

SOC from other sources (t/d)

2.3

5.6

4.0

6.8

2.3

1.1

22.1

Observed carbonaceous

BOD (t/d)

4.7

24.8

3.2

7.4

3.4

1.5

45.0

39.9

Degradable SOC (t/d)

4.1

21.5

2.8

6.4

2.9

1.3

38.9

34.5

Stable SOC

( t /d )

1.8

3.0

2.6

4.7

3.9

1.2

17.3

9.8

In its routine programme of water quality surveillance, the Trent River Authority (now part of the Severn-Trent Water Authority) measured both the total BOD of river waters and their BOD when nitrification is suppressed by the addition of N-allylthiourea. The results of the latter test are commonly referred to as the carbonaceous BOD (B) and should bear some relationship to the degradable organic-carbon contents of river waters.

If, during the course of the 5-day BOD test, the river water behaves biochemically in the bottle as it does in travelling through the river, then the carbonaceous BOD content of the sample could be related to the degradable SOC content by the relationship B = N(1 - Y)C[1 - exp(- 5k)], where k is the decay coefficient deducible from a mass balance of carbonaceous BOD, N the oxygen equivalent of the amount of SOC which is oxidized to CO2 in five days, Y is the yield coefficient for the conversion of SOC to bacterial cells, and C is the initial degradable SOC content. The value of the oxygen equivalent (N) could, in principle, be derived from chemical oxygen demand tests, and will vary with the oxidation state of the soluble organic carbon. For example N would have the values 4, 3.33, and 2.67 parts by weight of 0 2 per part of carbon in oxidation states of -2, -1, and 0, respectively.

Solution of the mass balance equation B = Zbexp(- kd), for loads of carbonaceous BOD given in Table 2, where b is the input mass flow from a tributary located a distance d above the point where the mass flow (B) is measured in the Lower Trent, yields a value of 0.00249/km for the decay coefficient k which is equivalent to a value of 0.137/d at an average velocity of flow of 55 km/d. Hence the decay factor [1 - exp(- 5k)] equals 0.496 and the half-life of degradable SOC is about 5 days. Taking a value of Y of 0.3 the relationship between carbonaceous BOD and the degradable SOC content becomes B = 0.35NC. It is unlikely that the carbon in the majority of organic compounds derived from effluents and natural sources will


be in an oxidation state lower than -2. For example, only saturated hydrocarbons display oxidation states for carbon less than -2, e.g. methane (-4), ethane (-3), propane (-2.67), and C H + (approaching -2 as n grows larger). Accordingly, and arbitrarily, an oxidation state of -1 is assigned to the degradable portion of the soluble carbon, and the relationship becomes C = 0.865B. Values of the mass flows of degradable carbon calculated on this basis are given in Table 2. Regression of values of the loads of SOC on the loads of carbonaceous BOD (B) using as a model the equation C = S + aB where S is the non-degradable portion of the SOC, and a is a constant, yields estimates for S of 2.63 t/d for each tributary (total flux 15.8 t/d) and a value for a of 0.899. Since a is related to N by the equation 1/a = N(1 - Y)[1 - exp(- 5k)], after insertion of values of 0.3 for Y and 0.496 for [1 - exp(- 5k)] the value of N is found to be 3.2 parts 0 2 per part carbon, corresponding to an oxidation state of organic carbon slightly greater than the value of -1 assumed above.

Comparison of the calculated mass flows of degradable carbon with the values of the total flow of SOC listed in Table 2 for each tributary indicates that the major portion of the SOC load is degradable. Comparatively little removal takes place in the river, however, because the average travel times of waters from the various tributaries to Nottingham are much less than 5 days. (For example water from the Upper Trent takes on average about 1.3 days to reach Nottingham.)

Taking the mass flow reaching the Lower Trent to comprise 34.5 t/d of degradable organic carbon and 17.3 t/d of stable carbon, the average concentration of organic carbon in water abstracted from the Trent as a function of storage time (t) would be predictable, as a first approximation, from the equation C = 2.4 +4.9 exp(- 0.137t) where concentrations of 2.4 and 4.9 mg/1 correspond to the mass flows of 34.5 and 17.3 t/d, respectively. Thus after 7 days storage as recommended by Miller and Short (2) the concentration of SOC would be 4.4 mg/1. After 30 days the concentration would be close to the non-degradable residue (assuming this to be stable) and the problem of rendering potable water wholesome would then depend upon the efficiency with which this residue could be removed by treatment.

MINOR CONSTITUENTS OF THE ORGANIC LOAD

Minor constituents of the organic carbon load that were determined in 1970 are phenolic substances reported as phenol (C^H^OH) and anionic detergents reported as Manoxol OT. Residues of pesticides, reported as total pesticides, were determined after 1970 in between 1 and 3 samples per year. These occur at concentrations which are less than 0.5 μg/l and the corresponding mass flows from each tributary are less than 1 kg/d. The mass flow in the Lower Trent over the period 1971-75 could be reckoned to be about 1.5 kg/d and this comprises about 12.5 parts per million by weight of the soluble organic carbon load or 40 parts per million of the particulate carbon, if adsorbed.

Some difficulty attends the calculation of a mass balance for phenolic substances, since these are detected in less than half the samples collected. This results in average concentrations for the year being lower than the lowest concentration found, and a mass balance calculated on this basis would be unrepresentative of the behaviour of phenolic substances when they are present in the river water. Accordingly, the estimates of mass flow reflect a situation which might occur on about 1 in 3 occasions when river water samples are collected. Examination of the data suggests that when phenolic substances are detected in the river they occur mainly in the first and last quarters of the year, indicating that the mass flows may be substantially influenced by the run-off of natural products and/or by being destroyed in the river during the summer months.


Inspection of the results given in Table 3 indicates that the phenolic residues behave conservatively and occur at concentrations which are many times that of phenol itself, which was found by Miller and Short (2) to occur at concentrations between 0.001 and 0.016 mg/1, with an average of 0.003 mg/1, at Nottingham over the period 1970-71. These two observations would support the view that the flow of phenolic substances in the Trent comprises stable compounds released during the wetter months of the year, probably from road surfaces in the urban areas. The organic carbon equivalent of these substances when expressed as phenol is 76.6 per cent. Thus, of a tributary input of 56.2 t/d of soluble organic carbon, phenolic compounds comprise about 1 per cent.

TABLE 3 Mass Flows and Mean Concentrations of Phenolic Substances and Anionic Detergents

Tributary

Upper Trent

R. Tame

R. Dove

R. Derwent

R. Soar

R. Erewash

Total

Lower Trent

Phenolic substances


0.11

0.13

0.10

0.09

0.11

0.11

0.13

Mass flow (t/d)

0.12

0.25

0.13

0.14

0.09

0.02

0.75

0.87 ,

Anionic detergents


0.18

0.72

0.12

0.17

0.22

0.43

0.23

Mass flow ; (t/d)

0.20

0.34

0.15

0.29

0.19

0.06

2.23

1.56

In contrast to phenolic substances, anionic detergents appear to be degraded rapidly in the river system. Solution of the mass balance equation yields a decay coefficient of 0.008/km (0.44/d), indicating a half-life of about 1.6 days. The carbon content of Manoxol OT, in terms of which anionic detergents are reported, is 54.5 per cent. Hence the total tributary load of 1.23 t/d corresponds to an organic load of 1.2 t/d and comprises 2 per cent of the soluble carbon load.

NITROGENOUS AND CARBONACEOUS OXYGEN DEMANDS

It was deduced earlier that 5.1 t/d of the carbonaceous BOD load from the tributaries was destroyed in the river at a rate characterized by a first-order coefficient of 0.137/d. This corresponds to a daily consumption of 10.3 t/d of oxygen in the river system by the ultimate carbonaceous load (L0). The total tributary input of ammoniacal-N in 1970 amounted to 14.1 t/d of which only 7.1 t/d survived in the river at Nottingham. A loss of 7 t/d of ammoniacal-N would exert an oxygen demand upon nitrification of 30.3 t/d oxygen. The first-order decay coefficient for the oxidation of ammonia has a value of 0.8/d (half-life 0.87 days). Thus nitrification dominates the process of oxygen consumption in the self-purification phenomenon and is of great benefit in reducing the ammoniacal-N content of Trent water.


In studying the treatability of Trent water, Miller and Short (2) found that 80-100 per cent of the ammonia disappeared during 7 days of storage after a maturation period of about 12 weeks in the experimental storage basin. A decay coefficient of 0.8/d in the river is quite consistent with these observations. Firstly, it has been shown elsewhere (6) that ammonia is removed in the Trent by benthic nitrification, and a 12-week period of maturation would probably be the time necessary for a highly active deposit of nitrifying bacteria to become established. Secondly, the removal of ammoniacal-N at a rate given by the rate coefficient of 0.8/d would result in a 96.4 per cent loss in 7 days.

CONCLUSIONS

1. The load of organic carbon imposed upon the R. Trent by its tributaries comprises 72 per cent soluble organic carbon and 28 per cent particulate carbon. The major fraction of the soluble organic carbon is derived from effluents (61 per cent).

2. A relationship between the soluble organic carbon load and the load of carbonaceous BOD was derived and indicates that 68 per cent of the load of SOC may be degradable with a half-life of about 5 days and an effective oxygen demand, corresponding to an oxidation state of -1, of 3.33 parts 0 per part carbon.

3. Because of the slowness with which soluble organic carbon is apparently oxidized in river waters, substantial reduction by self-purification would only occur after about 30 days when a non-degradable residue of 32 per cent would remain.

4. Phenolic substances and anionic detergents comprise almost 3 per cent of the soluble organic carbon load. Phenolic substances appear to be conserved whereas anionic detergents appear to be rapidly degraded in the river.

5. Ammoniacal-N is removed from the river with a half-life of about 0.9 days. Half of the input load is nitrified and this process would consume 75 per cent of the oxygen demand exerted by self-purification in the river. Nitrification is of substantial benefit in reducing the ammoniacal-N content of the waters of the Lower Trent at a point where abstraction for public supply has been considered. The rate of nitrification in the river is quite consistent with losses of ammoniacal-N observed in water stored for 7 days after abstraction.

ACKNOWLEDGMENT

This paper is presented by permission of the Director, Water Research Centre.

REFERENCES

(1) Council of the European Communities. Council directive of 16th June 1975 concerning the quality of surface water intended for the abstraction of drinking water in the Member States, No. 75/440/EEC, Official Journal of the European Communities, L194/26 to L194/31. (25.7.1975).

(2) D. G. Miller, and C. S. Short. 3. Treatability of River Trent water. (1972). Symposium on Advanced Techniques in River Basin Management: the Trent Model Research Programme, The Institution of Water Engineers, London.


(3) Sir Norman Rowntree. Summary of Symposium. (1972). Symposium on Advanced Techniques in River Basin Management: the Trent Model Research Programme, The Institution of Water Engineers, London.

(4) J. H. N. Garland, and I. C. Hart. The effects of pollution on river quality. (1972). Report of the Trent Steering Committee Vol. IV. Her Majesty's Stationery Office, London.

(5) Water Quality Advisory Panel. (1976). Water Quality 1974/5. Severn-Trent Water Authority, Birmingham.

(6) J. H. N. Garland. Nitrification in rivers: studies in the Trent basin. Proceedings of a Symposium on the Use of Mathematical Models in Water Pollution Control. (Ed. A. James). 2, 15-54 (1973). Department of Civil Engineering, University of Newcastle upon Tyne.

The Scientific Aspects of the Chemical Substances Control Law in Japan

S. SASAKI

Ministry of International Trade and Industry, Japan

A. An Outline of the Chemical Substances Control Law in Japan.

The Chemical Substances Control Law has been effective since April 1974 in Japan, This law aims at preventing man from suffering health injury by way of en-vironmental pollution caused by the scattering and accumulation of harmful chemi-cals into the environment. Japan has suffered human injuries caused by PCBs {poly-chlorinated biphenyls). The new law makes provisions for excluding harmful new chem-icals like PCBs from arriving on the market, thus protecting man against health in-jury by such chemicals.

According to this law all new chemical substances must not have more than one of such hazardous properties as non-biodegradability persistence, high degree of bioaccumulation or chronic toxicity before the chemical substance can be produced or imported commercially. That is to say, a chemical which has found to have all three above-mentioned hazardous properties by the already available examination or newly tested data is prescribed as a specified chemical substance by government or-der. The specified chemical substance must be produced or imported at the minimum quantity necessary to meet the demand and only when its use cannot be substituted by other substances. As for the production of a specified chemical substance, the equipment for manufacturing it must conform to technical standards. Furthermore, no one can use a specified chemical substance for any use other than those pre-scribed by government order for each specified chemical substance. And no one can import any product prescribed in which a specified chemical substance in used. Now PCBs are prescribed as specified chemical substances. A diagram of the procedures of the law is shown in Fig. 1.

B. The test of new or existing chemicals

(1) The outline

In addition to the above-mentioned examinations and regulations concerning new chemicals, the government has been carrying out examinations on biodegradabilities of about 300 existing chemical substances, bioaccumulations of about 100 such chemi-cals and chronic toxicities of 19 such chemicals from 1974 to 1977.

The tests have been camident in an authorized government institute, The Chemi-cals Inspecting Institute, etc., by the same methods as those mentioned before for new chemical substances prescribed by the order of the Prime Minister, the Minister of Health and Welfare and the Minister of International Trade and Industry. From the results of the biodegradation tests of the bioaccumulation tests, 155 existing

283

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the Control Law in Japan.

Scientific Aspects of the Chemical Substances Control Law 285

chemicals were found to be safe and four chemicals (PCBs, hexachlorobenzene, poly-chloronaphthalene, mercuric oxide) to be poorly biodegradable and highly bioaccumu-lative. The final results of the chronic toxicity tests of existing chemicals in-cluding the latter three have not been obtained yet.

(2) The conditions of the biodegradation test

1. The ration of test substance to active sludge is usually 100 ppm: 30 ppm.

2. The test temperature is 25-2°C.

3. The pH of the supernatant of active sludge is 7.0(-l).

4. The test period is usually 2 weeks.

5. The activity of the sludge is confirmed by using aniline as a test substance. The percentage biodegradation of aniline calculated from the oxygen consump-tion value must exceed 40%.

6. The percentage biodegradation from the oxygen comsumption is calculated by the following equation:

ROD - R Percentage biodegradation = — — x 100(%)

BOD: Biological oxygen demand of the test substance (experimental)(mg)

B : Oxygen consumption of basal culture medium to which the inoculum is added (experimental) (mg)

TOD: Theoretical oxygen demand required when the test substance is completely oxidized (theoretical) (mg)

7. The percentage biodegradation from the result of direct analysis is calcul-ated by the following equation:

Percentage biodegradation = — ^ τ - — x 100(%)

Sa: Residual amount of the test substance or the residual amount of the total organic carbon after completion of the biodegradebility test (mg)

Sb: Residual amount of the test substance or residual amount of the total organic carbon in the blank test with water to which only the test sub-stance has been added (experimental) (mg)

The chemical analyses are carried out with the aid of total organic carbon an-alyser, gas chromatography, spectrophotometer, mass spectrometer and atomic absorp-tion apparatus for direct assay.

(3) The criteria for judging biodegradability

If the percentage biodegradation from the oxygen consumption exceeds 30% after 2 weeks from the beginning of the test and the result of a direct analysis is at least this value, the test substance is usually judged as well-biodegradable.

(4) The conditions of the bioaccumulation test

1. The test fish carp (Cyprinus carpio) 10 cm long and 30 g weight.

2. Before beginning the bioaccumulation test 48 hours TLm concentration of the test substance must be estimated by rearing mature orange-red

286 S. Sasaki

killifish (Ovizias latipes) in an aqueous solution of the test substance at different concentration levels.

3. The bioaccumulation test is usually carried out by rearing several carp in each of three suitable fish ponds, in two of which the aqueous solu-tion of the test compound is supplied successively at different concen-tration levels separately as low as possible within the analytical limit among the levels of 1/100, 1/1000 and 1/10,000 of 48 hours TLm concen-tration and in one of which there is pure water for the blank test. The apparatus is shown on Fig. 2.

Fig. 2. The bioaccumulation apparatus.

4. If the test substance has low water solubility suitable solubilizers (such as ethyl alcohol, fatty acid ester of polyoxyethylene sorbitan, etc.) are used.

5. The test water is kept at 25±2°C and the concentration of dissolved oxygen is kept at about 7 ppm.

6. The text is usually continues for 8 weeks and every two or three test fish are taken out not less than 10 hours after feeding and subjected to analysis.

7. The degree of accumulation, expressed by the concentration factor, is calculated by the following equation:

F FB „_ n -

CF : Concentration factor after n weeks n


FN : Concentration of the test substance in the fish body after n weeks in the test period

FB : Arithmetical mean concentration of the test substance in the fish body at the start of, and at the termination of, the blank test.

W : Mean concentration of the test compound in the aquarium.

5. The criteria for judging bioaccumulation

If CFnTs increase in proportion to n number and CF 8 weeks exceeds a few hun-dred times, the test substance is usually judged as highly bioaccumulative. In such cases the carp are kept in water for 2 more weeks and the excretion affect studied by measuring the speed of the decrease of the chemical.

Further, if necessary, the concentration factors in some important organs or tissues of the fish are measured separately and the critical effects looked at more carefully.

6. The conditions and criteria of the chronic toxicity test

The chronic toxicity tests of some existing chemical substances are practiced by almost the same methods as those of medicines in Japan. The test items are tera-togenicity9 tumorigenicity and pharmacokineticSj etc.

Table 1 shows the results of the biodegradation and bioaccumulation tests of the existing substances used over 100 tons a year in Japan and hazardous ones used in smaller quantities under the above mentioned conditions. Table 2 lists the new substances officially published as safe.

C. The relations of the major tendencies on biodegradability of tested chemicals with their structures.

(1) High biodegradabilities of straight-chained hydrocarbons (carbon number 2-16) have been ascertained.

(2) If a hydrocarbon has some tertiary united carbons, it becomes less biodegrad-able in proportion to the increase of tertiary united carbons.

(3) Benzene is said to be well-biodegradable, but it could not tested by the closed system oxygen consumption measuring apparatus because of its volatility.

(4) Aniline was so well-biodegradable that it is used as the standard for approving the activity of the sludge.

(5) Chlorinated aromatics or nitroaromatics were not almost biodegradable. But nitro-aromatics are expected to be degraded by some anaerobes.

D. The relations of the major tendencies on bioaccumulativity of tested chemicals with their structures.

(1) Chlorinated aromatics are rather highly bioaccumulative.

(2) As for chlorinated aromatics, the bioaccumulativities increase in proportion to the increased numbers of chlorine.

(3) PCBs are very highly bioaccumulative and the degree of accumulativity was shown by our tests to increase in proportion to the chlorine numbers, 6, 3, 2 and 4 in turn. In fact dichlorobiphenyl was 600-16,000 times at 2e2 ppb, 1120-10,300 times at 6.6 ppb during 8 weeks, trichlorinated-one was 5500-15,900 times at 1.8 ppb, 5900-20,200 times at 5.4 ppb during 6 or 5 weeks, tetrachlorinated-one was 5100-19,800 times at 2.1 ppb, 8100-21,900 times at 5.4 ppb, hexachlorinated-

288 S. Sasaki

one was 2500-9400 times at 14 ppb, 1700-7700 times at 42 ppb both during 6 weeks. These effects were not sufficient to the newly clarified effects that 3,4-vacant-one does not accumulate; Furthermore the rather high bioaccumulat-ivity of hexachlorinated-one comparing Cl numbers was because of the high con-centation of the solution for ease of analysis.

Here the bioaccumulated and solved polychlorobiphenyls were extracted by a n-hexane/ethanol mixture and detected by ECD gas chromatography,.

(4) Triohlorobenzenes (1,3,5 - 1,2,4 and 1,2,3) were moderately bioaccumulative. The degree of bioaccumulativity of 1,3,5, chlorinated benzene was higher than the other two. The degree of bioaccumulativity of the former one was 150 - 1620 times and those of others were 120-1320 times and 130-1200 times during 6 weeks.

(5) Eexdbromobenzene was not almost bioaccumulative.

(6) As for polymers like polythylene, polystyrene, Polyvinylchloride, polyphenyl-solicone and polymethyl methacrylate, the bioaccumulativities decrease in pro-portion to the increase of molecular weight (the degree of polymerization).

E. One way of thinking about the bioaccumulativity of polymers.

Here I mention the procedures, the effects and the way of thinking about the bioaccumulativity of some thermoplastic polymers. The tests were carried out in the Chemicals Inspecting Institute, etc., using styrene oligoiner, 14C-lab-elled polystyrene, vinylchloride oligomer and polymer, phenylsilicone oligomer and polymer and methylmethacrylate oligomer. From these tests we expected to verify that the high polymers are low bioaccumulative because of their usual water insolubilities. In order to practice the bioaccumulation test we had to solve the problem that polymers which have some proper solvent are easily dis-persed in water, but polymers which haven't any proper solvent are difficult to disperse in water. We tried to solve the problem applying the general rule that polymers of high degree of polymerization were easier to disperse in water and would provide more proper solvent than those of a lower degree if polymerization of the same chemical composition. For this purpose we, for example by ultrasonic wave radiation, by dielectrical heating, etc. We found that in order to under-stand the bioaccumulativity of high polymers we had to carry out tests using the polymers of a lower degree of polymerization or the oligomers. From the re-sults we were aware of the extent of the degree of polymerization within which the bioaccumulativity didnft change so much and the mechanism of bioaccumulation in fish. Now we come to the testing procedure of styrene oligomer and li+C-lab-elled polystyrene, vinyl chloride oligomer and polyvinyl chloride.

For the styrene oligomer we used the butyl (styrene) nH (n=l-14) structural one on the market, the butyl (styrene)n butyl (n=l-5) structural one and the butyl (styrene) nH (n=l-13) structural one both synthesized. The 48 hours TLm concentra-tion of the above three substances were over 1000 ppm, so the concentration levels of the rearing water were settled at 5 ppm for the first oligomer, at 2 ppm and 0.2 ppm for the latter two, using hydrogenated castor oil by 5 times in THF and evapor-ating THF. The analyses of the substances in the water were by gas chromatography after extraction by chloroform, dehydration by Na2S0i+ concentrated and measured to a definite volume. The analyses of the substances in the fish were by gas chroma-tography after extraction by n-hexane, cleansing by IN NaOH and water (only the first oligomer), dehydrated by Na2S0i+, concentrated and measured to a definite vol-ume.

Here the n-hexane extraction of the first ooligomer was separated by GPC and analysed by G.C.-M.S. The measured concentration levels of the water were maintained within a 20% deviation from the settled one, except that the lower concentration levels of the latter two oligomers changed up to 50% deviations. The concentration


factors were 12-293 times (n=2 fraction of the first oligomer), within 8.2 times (n=3 fraction), within 5,3 times (n 4 parts), 22-346 times (n=2 fraction of the second oligomer), within 11.6 times (n=3 fraction), 45-5529 times (n=l fraction of the third oligomer), 1.6-46.8 times (n=2 fraction) and within 2.6 times (n=3 frac-tion) during 4 weeks. From the fine analysis of every part of the fish body we found that some particles were apt to be taken into the intestine. As an effect the concentration factors of all these substances changed rapidly between the molecular weight 250-300.

As to the vinyl chloride oligomer (the mean degree of polymerization was 6) and the polyvinyl chloride (the mean degree of polymerization was 100) we also car-ried out the bioaccumulation tests. In these tests the concentration factor reached a maximum (15,100 times) at about 350 molecular weight and decreased rapidly over 500 molecular weight.

From these effects we studied the relation of the concentration factors of polystyrene and Polyvinylchloride measured by the above tests and the calculated oneb according to W.B. Neely's equation using the partition coefficients obtained under A. Lee et al. Ts method. The logarithmic calculated concentration factor curves to the logarithms of the partition coefficients of both the polystyrene (Bu-Bu type) and the Polyvinylchloride increase linearly, but the measured one of the polystyrene decreased linearly and the measured one of the Polyvinylchloride had its maximum (see Figs. 3 and 4). On the other hand a result of the bioaccumulation test of the

O Measured

Δ Calculated

4 6 β 10

T h e logarithm of par t i t ion coef f ic ient

Fig. 3. The relation of the concentration factors and the partition coefficients about styrene oligomer.

290 S. Sasaki

2 4 6 8 10 12

The logarithm of partition coefficient

Fig. 4. The relation of the concentration factors and the partition coefficients about styrene oligomer (Bu-Bu).

phenylsilicone oligomer (the molecular weight was 546), the phenylsilicone (the molecular weight was 12,000) and the methylmethaorylate oligomers (the means mole-cular weights were 169 and 1300), these substances were all found to be low bio-accumulative.

From this we can conclude that the ease of being taken into the fish body de-pends on the particle size as shown in the case of the vinyl chloride oligomer. Nevertheless the degree of fat solubility increases and the concentration factor changes to a decrease at a definite particle size. Moreover, like the methylmetha-crylate oligomer of the lower degree of polymerization, in the case of the substance being taken into a fish body and excreted after becoming water soluble by the clea-vage of the ester bond, the substance shows a low bioaccumulativity.

Furthermore, we studied the bioaccumulation mechanism using ^C-polystyrene of 6.19x 104 mean molecular weight. The sample was produced by the NEN company. It had a 0.495 m Ci/mole specific activity and the radiochemical purity was 99%. We pre-pared 11+C-polystyrene dispersedliquid by the same method as above. In a 5.56 ppm solution we reared carp, not giving any bait for 7 days. The picked carp was hole-body-au-toradiographed. Only in the intestine were small black shadows ascertained. In addition to this inspection the carp were picked up after 1 hour, 4 hours, 12 hours, 3 days and 7 days from the beginning and the radioactivity of every organ or issue was measured at those times. The mean concentration level in the water was


2.96 ppm (31,700 dpm/ml) for the rearing period. The concentration factor went up slowly as the time passed, but the absolute values were under 3.66 times. We also injected 12xl03 Ci/g body weight ll+C-polystyrene prepared by solving in THF and ev-aporating THF into the carp muscle from the back gradually. After the injection the carp was reared for 3 days in water and hole-body-autoradiographed. In the radiogram the injected part was clearly blackened and the kidney, the spleen and the intestine were slightly blackened. Further the carp was forced to swallow the bait containing the 1I+C polystyrene treated by THF in a quantity of about 7xl06 dpm/1 carp. After the carp was reared in water and given no bait for 3 days, the carp was hole-body-autoradiographed. Only the intestine and the gall were slightly blackened. These effects were ascertained more finely by measuring the radioactivity in every organ and tissue using a liquid scintillation counter. We can conclude that a small quan-tity of the ll+C-polystyrene can be taken rapidly into the intestine and only a lit-tle of it can be distributed to the whole body very slowly.

F. Conclusion

Japan will be continuing the biodegradation tests and bioaccumulations tests of new or existing chemical substances on a large scale.

The results of these tests will be published in official reports and will give us useful information.

TABLE 1

1. Substances confirmed to be well-biodegradable

(1) Low molecular chain-like organic compounds

Dime thy lamine

Lauryldinethylamine Hexamethylenediamine Allyl alcohol iso-Butanol sec-Butanol 2-Methyl-2, 4-pentanediol Epichlorohydrin N,N,N-trimethyl-2-hydroxyethylammonium chloride Diethanolamine Monoethanolamine 2-n-Butoxyethanol 3-Methoxy-n-butanol N,N-dimethylethanolamine Methylisobutyl ketone 4-Methyl-3-pentan-2-one Acrylic acid Acrylic acid amide Methyl aerylate Ethyl acrylate Butyl acrylate Octyl acrylate 2-Ethoxyethyl acetate 4-Methoxy-n-butyl acetate Methylmethacrylate Acetic acid monochloride Formamide Acetonitryl Isobutylnitryl

292 S. Sasaki

α-Chlorofumalic acid diethyl ether Fatty acid epoxide glycerine ester o,m,p-Tricresyl phosphate Na salt of higher alcohol sulfuric acid ester Polyoxyethylene Lead stearate Na salt of sulfuric acid ester of polyoxyethylenemonoalkylether n-P en t ad e c ane

(2) Low molecular carbo-monocyclic organic compounds

Terephthalic acid Cyclohexanone Benzyl chloride Acetanilide Hydroquinone Isophthalic acid Resorcine Phthalic acid unhydride Benzonitrile Salicyclic acid 2-Me thoxyaniline 4-Methoxyaniline N-acetoacetyl-4-methylaniline Isopropyl benzyl alcohol Acetophenone p-Dichlorobenzene Xylene Dibutyl phthalate Diheptyl phthalate Dioctyl phthalate Diisodecyl phthalate Cresol Xylenol Isopropyl benzene Dicyclohexyl phthalate Na salt of alkylbenzensulfonic acid (soft type) 2-Me thy1aniline 4-Me thy1aniline N-acetoacetyl-2 methylaniline 4-Aminobenzene-l-sulfonic acid Alkylbenzene (straight chain type)

(3) Other type compiunds

Biphenyl Furfural Piperidine 3-Acetyloxoran-2-one Benzo a-pyrone 2-Me thyIpyridine Pyridine Tetrahyrofurane a-Naphthol Phosphoric acid diphenyl monoorthoxenyl Styryl xylene


2. Substances confirmed to be not almost biodegradable

(1) Inorganic compound

Mercuric oxide

(2) Low molecular chain-like organic compounds

Dibromoethane

Tetrachloroethylene Tetrabromoethane Dinitrobenzene 4-Nitrophenylmethyl ether l-Amino-2-methoxy-5-methylbenzene 4-Methylphenylene-l,3-diamine Dibromocrezylglycidyl ether Hexab romob enz ene Chlorocyclohexane 2,5-Dichloroaniline 3,4-Dichloroaniline Chloromethyl aniline 4-Chloro-2-nitroaliline p-Octyl phenol 4-tert-Butyl-hydroxyphenol N-acetyl-4-ethoxyaniline o-Nitrophenol o-Chloro benzo ic acid Na salt m-nitrobenzenes'ulfonate Na salt chloro nitro benzenesulfonate p-Chloro toluene 2,5-Dimethylaniline Polyoxyethylenealkylphenyl ether p-tert-Butylphenol o-Chloronitrobenzene 3-Aminopheno1 4-Ethoxyanil ine o-Chlorotoluene 1,3-Dicyanobenzene 3,4-Dimethy1aniline p-Chloronitrobenzene

(4) Other type compounds

Polychloro diaminodiphenylmethane Dodecachloro-dodecahydro-dimethanodibenzocyclooctene Polychlorinated triphenyl Polychlorinated biphenyl Polychloronaphtalene Melamine 2,2-bis (4-Hydroxyphenyl) propane Anthracene 2-(N-phenylamino) naphthalene l-Phenylamino-4-isopropylaminobenzene 1,3,5-Trichloroisocyanuric acid 2-Naphthylamine sulfonic acid 1,4-Dioxane

294 S. Sasaki

2,2-bis (4-Hydroxy-3,5-dibromopnenyl) propane Ν,Ν-Diphenyl guanidine 7-Amino-4-hydroxy-2-naphthalenesulfonic acid K salt 7-hydroxy-l,3-naphthalenedisulfonate Na naphthionate Na salt 4-amino-5-hydroxy-2,l-naphthalenedisulfonate Piperazine Decabromodiphenyl Decabromodiphenyl ether Na salt l-naphthol-3,6-disulfonate Na salt 2-naphthol-4-sulfonate 2-Amino-8-naphthol-6-sulfonic acid 6-6f-Ureyl-bis (Na salt-naphthol-sulfonate) 1-Chloro anthraquinone 2-Chloro anthraquinone Quinoline 2-Mercapto benezoimidazole Carbanole Terphenyl N-phenyl-1-naphthylamine 5-Amino-4-naphthylsulfonic acid 1,3,5-tris (2-Hydroxyethyl)isoeyanuric acid 2-Chloro-4,6-bis (ethylamino) -s-triazine 2-Chloro pyridine 2-(2-Hydroxy-3,5-di-tert-butylphenyl)-5-fruoro benzotriazole N,N-diphenyl-p-phenylenediamine Dichloro-1,4-naphthoquinone mono di or tri-(a-Methylbenzyl) phenol 2,4,6-Trichlorophenol-(4l-nitrophenol) ether

3. Substances confirmed to be nonaccumulative or low accumulative

Dodecachloro-dodecahydro-dimethanodibenzocyclooctene Tetrachloroethylene Monochlorobenzene Dinitrotoluene Diheptyl phthalate Dibutyl phthalate Dioctyl phthalate Diisodecyl phthalate D ime thy1fo rmami d e Nitrobenzene Pentaerythritol Nonylphenol p-Chloronitrobenzene p-tert-Butylphenol o-Chloronitrobenzene o-Nitrotoluene 2-(N-phenylamino) naphthalene o-Toluenesulfonamide p-Nitrotoluene 2-Naphtylamine sulfonic acid o-Nitroanisol Ν,Ν-dimethylaniline 1,4-Dioxane Alkylbenzene (branch type) Polyoxypropylene 1,5,5-Trimethyl-2-cyclohexane-3-one

Scientific Aspects of the Chemical Substances Control Law

Ν,Ν-diphenyl guanidine 2-Mercapto benzothiazole o-Chloroaniline 2-Amino-5-chloro-4-methylbenzene sulfonic acid Sodium naphthionate p-Nitrophenol tert-Butyl alcohol Chlorinated paraffin 2,2-Dichlorodiethyl ether N-me thy1aniline N-ethylaniline 2-Nitroaniline 3-Nitroaniline 4-Nitroaniline 2,2,4,4,6,8,8-Heptamethylnonane n-Pentadecane Carbazole Hexabromobenzene 4-Nitro phenylmethyl ether (p-Niroanisol) Tetrahydrothiophene-1,1-dioxide l-Amino-2-methoxy-5-methylbenzene

Substances confirmed to be moderately accumulative

o-Dichlorobenzene 1,2,4-Trimethylbenzene Diphenylamine 2,2-Bis (4-hydroxy-3,5T-dibromophenyl) propane Trichlorobenzene

4. Substances confirmed to be highly accumulative

Polychlorinated biphenyl (Cl:2,3,4,6) Hexachlorobenzene Polychloranaphthalene Mercuric oxide

TABLE 2

List of New Substances Officially Published to be Safe

( ) : Biodegradable [ ]: Nontoxic No remark: Nonaccumulative or low accumulative

(1) Hexamethylene-bis [ hydroxyfatty acid (c=16-18) amide] (2) 2-(l-Hydroxyethyl) acrylonitrile (3) Sodium hydroxyacetate (4) Hydrolysis products of polya-ljo'-D-galacto-ß-l^-'-D-mannose (1:2-4) (5) Reaction products of starch and magnesium chloride (6) Chelate compounds of lignin sulfonic iron-magnesium-copper-zinc-molybdenum (7) Diethylaluminium monomethoxide 8 9-Eicosyl-9-phosphobicyclo [3,3,1 ] nonane 9 9-Eicosyl-9-phosphobicyclo [4,2,1 ] nonane

[10] 3-Acetyl-6-methyl-2H-pyrane-2,4(3H)-dione (Zn salt) (another name: dehy-droacetate, Zn salt)

11 2-Hydroxymethyl-3-methylbutane-l,3-diol

296 S. Sasaki

12 3-Methylpentane-l,3,5-triol 13 3-Methylpentane-l,5-diol 14 2-Hydroxy-5-n-nonylbenxophenone oxime 15 1,3-Diamino-2,4,6-trinitrobenzene 16 Cyclopentadiene-isoprene-acrylnitrile diene reaction products, hydrogenated

and hydrolyzed (potassium salts) (17) Methyl 3-mercaptopropionate 18 7-Hydroxy-5,8-diethyl-6-dodecanone oxime (19) Cyclopentene (20) Ethylene cyanhydrin (21) Dimethyl l-methyl-2-(l-phenylethoxycarbonyl) vinyl phosphate (22) Sulfonated natural fatty acid, methyl ester (Na salt) (23) Sulfonated natural fatty acid, ethyl ester (Na salt) 24 o-Chlorostyrene 25 m-Chlorostyrene 26 p-Chlorostyrene 27 Zinc phosphate 28 Disodium 5-acetylamino-3-[ 4-{2-(2f-N-n-dodecylcarbamoyl-lT—ydroxynaphtho-

4f-yloxy)-ethoxy}-phenylazo] -4-hydroxy-2,7-naphthalene disulfonate 29 l-Benzyl-5-ethoxy-3-[ {N-(2-chloro-5-{N-{2-(2,4-di-tert-pentylphenoxy)-

butanoyl}-amino}-phenyl}-carbamoyl}-(4-methoxybenzoyl)methyl}-hydantoin (30) N-(Phosphonomethyl)-glycine (31) Sucrose benzoate 32 2,5-8-Trimethyl-2,5-8-triazanonylstyrene-styrene-divinylbenzene copolymer 33 Dicyclopentadiene-vinyl acetate copolymer 34 Ethyl N-methyl-N-phenyl-carbamate (35) n-Butyl 2-(benzoyloxy)-propionate 36 Methoxybenzenediazonium hexafluorophosphate 37 2-Hydroxyethylmannan 38 Carboxymethylmannan (Na salt) (39) 2-Methyl-4-isothiazolin-3-one (40) 3-Cyclohexyl-6-(dimethylamino)-l-methyl-l,3,5-triazine-2,4(lH,3H)-dione 41 Styrene divinylbenzene-acrylonitrile copolymer (Na salt) (42) l,lf-(l,4-Phenylene dicarbonyl)bis(hexahydro-2-azepinon) (43) 4-Hydroxyphenyl-a-ketoacetohydroximic acid chloride 44 3,5,5-trimethyl-l-p-methoxycarbonyl-phenoxycarbonyl amino-3-p-methoxycarbonyl-

phenoxycarbonylaminomethyl cyclojiexane 45 Perfluoro (n-propylvinyl ether) (46) Formadehyde condensation products of protein hydrolysis products and their

sodium salts 47 4-Methoxydiohenyl amine (48) N,NT-bis(l,3-dimethylbutylidene)-hexamethylene diamene (49) 9,10-Epoxy-oleyl acrylate (50) 9,10-Epoxy-oleyl methacrylate (51) l,4-Dihydro-9,10-anthracene dione (52) Disodium salt of l,4-dihydro-9,10-anthracene diol 53 l-(5-Tert-butyl-l,3,4-thiadiazol-2-yl)-3-methyl-5-hydroxy-2-imidazolidionone 54 Tetrafluoroethylene-ethylene-propylene copolymer 55 (2-Hydroxy-5-tert-nonyl)acetophenoneoxim (56) Hexamethylene-bis-(4-hydroxybenzoate) 57 Maleic anhydride denaturation products of ethylene-vinyl acetate-copolymer 58 2,4-Dinitro-5-[ 2-methyl-4-(N,N-di-2-acetoxyethylamino)phenylazo] thiopene (59) N-Methylacetamide 60 Dimethyl-bis(N-methylacetamido) silane 61 1-Alkene (c=20-28)·Ν-[ NlNf-dimethyl-Nl-(2f,3f-epoxypropyl)-ammoniopropyl)

maleimide chloride copolymer (62) Bis(3,4-epoxycyclohexylmethyl) adipate (63) Di-n-octyl ether


(64) Di-n-decyl ether (65) Di-n-dodecyl ether 66 3-{2-Chloro-4-(n-tetradecanoylamino) anilino}-4-{4-(2,2-dimethyl propanoy-

lamino) phenylazo}-1-(2,4,6-trichlorophenyl)-5-oxo-2-pyrazoline 67 2-Ethylhexyl-3,4-methylenedioxy-a-cyanocinnamate 68 Sodium salt of bis [ a,x,w-trihydro-perfluoroalkyl (c=3,5) ] sulfosuccinate (69) 2-Hydroxy-2-(2-hydroxyethylamino-carbonylmethyl) succinic acid 70 3-Methyl-3-methoxybutanol (71) 2,2f-Oxydiacetate 72 Poly(triallylamine sulfate) 73 Cross linkage reaction products of polyvinylalcohol and glutalaldehyde (74) 1,3-Diphenyl-l,3-propanedione 7 5 1-(5-Tert-buthy1-1,3,4-thiadiazol-2-y1)-1,3-dimethy1 urea 76 Sodium salt of O-0-di-isopropylthiolphosphate (77) Zinc p-toluene-sulfonate (78) Ether bond substances of cellulose-ethyleneoxide reaction products and I-

(N,N,N-trimethylammonio)-2,3-epoxypropane chloride 79 a-Pivaloyl-a-[ 5 or 6-(3-methyl-2-benzothiazolinylidene)amino-l-benzotriaz-

olyl] -2-chloro-54 l-(2,4-di-tert-pentylphenoxy)-buthylamido] acetanilide 80 Sodium 4-brome-2,5-dichlorophenolate 81 2,4-Dinitro-5-[4-[N,N-bis(2-acetoxyethyl)amino] phenylazo] thiopene 82 5-[2-(2-carboxyphenylazo)acetoacetylamino ]benzimidazolone (83) Hexadecanedioic acid (84) Eicosane dioic acid (85) l,4,4a,9a-Tetrahydro-9,10-anthracenediol (86) l,4-Dihydro-9,10-anthracenediol (87) 2-n-Octyl-4-isothiazolin-3-one 88 3-Phenoxybenzylcyanide 89 2-(2,2-Dichlorovinyl)-3,3-dimethylcyclopropane-carboxylic acid 90 Perfluoro copolyalkylene (propylene and methylene)polyether (91) Methyl-2-chloropropionate 92 N-ethyl-N-2-(methylsulfoneamido) ethyl-3-methyl-4-nitroaniline (93) 3-(4-Hydroxylphenyl) propionic acid (94) Di (4-Methylphenyl) carbonate (95) 2, 2f-{[0xybis (methylene)] bis[2-(hydroxymethyl)-l, 3-propanediol ester

of lower fatty acid (C=2-8) 96 4,4f-[2,2,2-Trifluoro-l-(trifluoromethyl) ethylidenel bisphenol (97) l-Amino-2-phenylazo-7-[5"-(2,lf - chloro-4tfl -aminotriazine-6fff-yl)-

aminophenylazo ] -8-hydroxynaphthalane-3,6,2f, 5f, 2fl-pentasulf onate (98) n-Dodecyl-3, 4-methylenedioxy-a-cyanocinnamate (99) Dicyclopentadiene . allylalcohol copol3rmer 100 1, 2-bis (2,4,6-Tribromophenoxy) ethane (101) Di-n-propylamine (102) Epoxidated (oleyl oleate) 103 Sodium 3-[[2,5-dichloro-4-[(ethylamino)sulfonyl ]phenyl ]azo]-2-methyl-LH-

indole sulfonate (104) Diethyl glutalate (105) Di (2,3 - epoxypropyl) 8 - hexa-decenedioate (106) Di (2,3 - epoxypropyl) 8, 12 -eicosanedienedioate (107) Di (2,3 - epoxypropyl) 7 - vinyl-tetradecanedioate (108) Di (2,3 - epoxypropyl) 12 - vinyl-8-octadecenedioate (109) 8-Hexadecenedioate (110) 8, 12-Eicosadienedioate (111) 7-Vinyltetradecanedioate (112) 12-Vinyl-8-octadecenedioate (113) Di(2,3-epoxypropyl) 7-ethyl-tetradecanedioate (114) Di(2,3-epoxypropyl) 7-ethyl-octadecanedioate (115) 2,5 - Biphenyldiol

298 S. Sasaki

116 Propylene · 1-butene . ethylene copolymer (117) Nitrilotrisulfonic acid (118) Ammonium nitrolotrisulfonate (119) Triisostearoyloxy-isopropoxytitanium 120 Terephthalic acid - 1,4 - butanediol - 2,2* - isopropylidenebis (dibromo-

p-phenyleneoxy) diethanol copolymer 121 4,4* - Methylenebis (2,6-dibromoaniline) 122 N-(3,5-dichlorophenyl)-l,2-dimethylcyclopropane-l,2-dicarboxy imide (123) 2-Imidazolidone.glyoxal condensate 124 4-(n-Octadecylsuccinimido)-2-(l-phenyl-5-tetrazolylthio)indan-l-on 125 4- [4-(8-Acetoamido-l-hydroxy - 3,6 - disulfonaphtylazo) phenoxy]-N-[4-(2,

4-di-tert-pentylphenoxy) butyl]-l-hydroxy-2-naphtho-amide disodium salt 126 2-(l-Allyloxyaminodutylodene)-5,5-dimethyl-4-methoxy-carbonylcyclohexane-

1,3-dion sodium salt (127) Potassium-tert-butoxyde (128) Butylethylamine 129 Polyalkyl (C=l - 4) (mercapto-propyl)polysiloxane (130) Methyleneaminoacetonitril trimer 131 Polybutadiene (terminal hydroxy group) hydride.maleic anhy-dride.epichloro-

hydrin polycondensate 132 4-(2-Bromo-6-chloro-4-nitrophenylazo)phenyl-3,3'-iminobis (methylpropionate) 133 3-Acetylamino-4-(4-methoxy-carbonylphenylazo)phenyl-3,3'-iminobis (methyl-

propionate) 134 3-Acetylamino-4-(4-nitro-phenylazo)phenyl-3,3'-iminobis (methylpropionate) 135 5-Acety1amino-4-(2,4-dinitro-phenylazo)-2-methoxylpheny1-3,3'-iminobis

(methylpropionate) 136 N-[2-(2,4-dinitro-5-thienylazo)-5-(N,N-diethylamino)phenyl ] acetamide 137 l,5-Dihydroxy-4,8-bis[2-{2-(methoxyethoxy)carbonyl}ethylamino ]anthtaquinone (138) Calcium chlorite 139 Poly(15-150)[{hydroxypoly(1-3) ethoxyethyl-poly(2-4)oxyethylene}borate] (140) Reduced keratin (141) N-methylacetoacetamide (142) 4-Hydroxy-2H-l-benzopiran-2-one (143) 2-Me thy1-2-propenylidene-diacetäte 144 l-Benzyl-5-ethoxyhydantoin (145) 2,2'-Oxydiethylbis (chloroformate) 146 1-Butene-propylene-copolymer 147 Ethylene-4-methyl-l-pentene copolymer 148 Cu complex salt of l-hydroxy-2-[2'-hydroxy-3'-(5n-chloro-2!l,4" difluoro-6"-

pyrimidylamino) phenylazo ] naphthalene 3,6,5'-trisulfonic acid NA salt 149 1:2 type Cr complex salt of l-Amino-2-(2'-hydroxy-3',5'-dinitro-l'-pheny-

lazo) naphthalene-4-sulfonic acid Na salt (150) Acrylic acid 2-[2-(ethyloxy) ethyloxy] ethyl 151 2-Bromo-3-5-dichloropyridine 152 2,3,5-Trichloropyridine 153 2-[4-(3,5-dichloro-2-pyridinyloxy)phenoxy] propionic acid 154 2-[4-(3,5-dichloro-2-pyridinyloxy)phenoxy ] propionic acid Na salt (155) Hexahydro-1,3,5-triethyl-l,3,5-triazine

Metabolism: Detoxification or Toxification S. SAFE, C. WYNDHAM, A. CRAWFORD and J. KOHLI

Guelph- Waterloo Centre for Graduate Work in Chemistry, Dept. of Chemistry, University of Guelph, Guelph, Ontario, N1G2W1, Canada

INTRODUCTION

The uptake or administration of a xenobiotic (i.e. foreign chemical) to an animal can result in a wide variety of responses. The chemical might induce an immediate toxic effect, the toxic response might be delayed or, with the case of a benign compound, there may not be any harmful cellular response. A major response to foreign chemicals is their biotransformation or metabolism into more polar meta-bolic products which can be excreted directly or converted into more water soluble conjugates and excreted. The metabolic processes can occur in diverse tissues and organs (skin, heart, lung) however the hepatic system plays the major role in the biotransformation of diverse chemicals such as drugs, industrial pollutants, pesti-cides, food additives and other foreign chemical substances (1).

The major source of enzymes associated with metabolic processes are bound to the membranes of the complex network of interconnected channels or endoplasmic reticulum which proliferate liver cells. Mechanical rupture of the cells and subsequent fractionation can result in the isolation of these packets of enzymes as a microsomal fraction. The microsomal enzymes are diverse and include hydro-xylases and other oxidases, O-demethylases, N-demethylases, deaminases, hydrases, hydrolases and transferases. A major enzyme system responsible for oxidative metabolism is aryl hydrocarbon hydroxylase (AHH) a microsomal mixed function oxidase enzyme which biohydroxylates a wide variety of hydrocarbon substrates including the polynuclear aromatic hydrocarbon (PAH) carcinogens. Microsomal AHH contains 2 protein components, cytochrome P-450 and NADPH cyctochrome-C reductase (a flavoprotein), a lipid component (phosphatidyl choline) and an absolute require-ment for NADPH and molecular oxygen (2,3). The overall oxidative electron trans-port process is summarized in Fig. 1. The lipophilic substrate, R-H, is converted

NADPH Cyt.-C Cyt. P-450 (Fe44"*") \ / (Fe*"*")

I RH

Fig. 1. Microsomal hydroxylase electon transport chain

299

300 S. Safe, et dl.

into a more polar R-OH metabolite which, on excretion, represents a cellular detoxification process.

METABOLISM OF CHLORINATED AROMATIC HYDROCARBONS BY AQUATIC ORGANISMS

Previous studies have confirmed that aquatic species detoxify xenobiotics using metabolic pathways comparable to the observed in more biologically active species (4,5). Marine vertebrates contain microsomes which possess an array of enzymes which are reported to be present at activites which are lower than those observed in mammals such as rats, rabbits and guinea pigs. Adamson has shown that the hepatic azo-reductase activity of various fish and amphibia was only 10-15% of that observed from rat hepatic microsomes and this comparative data was typical for that of several microsomal enzyme activities (6). The biohydroxylation of aromatics by fish and frog species has also been reported using biphenyl, acetanilide, aniline and benzo[a]pyrene. Preliminary in vivo experiments with Salmo gairdneri using isomeric chlorobiphenyls as substrates did not yield any isolable metabolic pro-ducts however the frog species (7), Rana pipiens, proved to be an excellent model system for n_ vivo metabolic studies (8). This paper will initially focus on recent results obtained for the metabolism of halogenated aromatic compounds which have been identified as contaminants in water and in aquatic organisms.

Polychlorinated Biphenyl (PCB) Metabolism

Administration of PCB substrates in corn oil to frogs followed by extraction and workup of the water containing the frog excreta gave several hydroxylated products. 4-Chlorobiphenyl was metabolized to yield a phenol, a catechol and a methylated catechol as shown in Fig. 2. Comparable studies have also shown the formation of

Cl Cl Cl Cl

OH OH OH ^X ^ CH3

Fig. 2. Metabolism of 4-chlorobiphenyl by the frog

other metabolites which are formed on hydroxylation of the chlorophenyl ring moiety (9). 4,4'-Dichlorobiphenyl was also hydroxylated to give 4,4f-dichloro-3-biphenylol and Aroclor 1254, a commerical PCB and a major environmental pollutant, was also converted into mono-, di- and trichlorobiphenylols as determined by mass spectral analysis. The results confirmed the preferential metabolism of the lower chlorinated PCB isomers in other mammalian systems (10). Results from our labor-atory have shown that 4-chlorobiphenyl is also metabolized by fish species and a thin-layer chromatogram of the lipid extractables from the water used to contain the fish clearly show a polar radiolabelled product which co-chromatographs with the phenolic metabolite. It has recently been shown that 2,2f,5,5'-tetrachloro-biphenyl is metabolized by Salmo gairdneri to yield polar phenolic metabolites and

Metabolism 301

^ ^ ^ ^

-+■ ci.

r ^ i ^

V -|-C1,0H + I -|"Cl2,0H + { -|~C13»0H

^

Aroclor 1254

dpm(3H)

50 i

40 J

30

X 10 J 20J

10 1

Fig. 3. Metabolism of Aroclor 1254 by the frog.

20 40 " 60 Rf

i=!= 80

o o Standards 4?-chloro-4-biphenylol 4-chlorobiphenyl

Fig. 4. Radio thin-layer chromatogram of 4-chlorobiphenyl metabolites (fish).

metabolite conjugates (11) although the precise structures of the metabolic pro-ducts were not determined.

Mechanistic studies using 4-chloro-4[2H]-biphenyl gave the hydroxy metabolite which contained 79% of the original deuterium indicating a 1,2-migration of this atom from the site of hydroxylation to the adjacent carbon atom (i.e. the NIH shift). This result confirms the intermediacy of an arene oxide and this pathway is typical of the AHH-mediated metabolic pathways (12,13).

2H(H)

OH

Fig. 5. 2 H shift in the metabolism of 4-chloro-4[2H]~biphenyl by the frog

302 S. Safe, et dl.

Chlorinated Benzene Metabolism

Chlorinated benzenes are widely distributed in the aquatic environment and previous studies have shown that these chemicals are metabolically degraded to yield hydro-xylated and conjugated metabolites. Our investigation with frogs (8) confirmed that isomeric chlorobenzenes are converted into diverse phenolic products including some dechlorination products. Of mechanistic interest was the formation of 2,3,4,6 -tetrachlorophenol from 1,2,3,4-tetrachlorobenzene which presumably forms via an NIH shift as shown.

Fig. 6. The metabolism of 1,2,3,4-tetrachlorobenzene by the frog

Chlorinated Naphthalene Metabolism

Polychlorinated naphthalenes (PCN) have also been identified in the aquatic envir-onment (i.e. aquatic sediments) and their metabolism by frogs and other species also yielded hydroxylated products. Again mechanistic studies with the frog con-firmed or supported an arene oxide intermediate since the migration of both Cl and 2H was observed.

X = C1,2H

Fig. 7. The NIH shift of chlorine and deuterium in chloronaphthalene metabolism

THE ROLE OF ARYL HYDROCARBON HYDROXYLASE IN METABOLISM OF CHLORINATED AROMATICS

The AHH enzyme system is known to occur in the hepatic microsomes of aquatic animals although it has been reported that the enzyme activity in these species is less than that observed in most terrestrial species (6). Recent data has indicated that trout and salmon AHH are capable converting hydrocarbons into diverse meta-bolites. Moreover trout liver microsomes oxidize the carcinogen benzo[a]pyrene at a rate 5 to 10 times or 15 to 30 times faster than male rat liver microsomes based on mg of microsomal protein and units of cytochrome P-450 respectively (13). Thus it is clear that both aquatic and terrestrial animals metabolize an array of chemicals, including many aquatic pollutants both in vivo and in vitro and that the hepatic system is an important component of this process. The metabolic response to chemicals represents, in part, a detoxification route.

Metabolism 303

However it has also been shown that microsomal-mediated metabolic processes are associated with chemical carcinogenesis (17) in which the formation of the meta-bolically-activated arene oxides can result in cellular interactions other than metabolic detoxification. The electron deficient carbon centers in arene oxides

HYDROCARBON AHH

NADPH

02

w DNA.RNA Protein

MACROMOLECULAR ADDUCTS (toxification, T)

METABOLITES (detoxification, D)

LOW MOLECULAR WT. ADDUCTS (detoxification, D)

Fig. 8. Cellular decompositions of arene oxides

can react with cellular nucleophiles such as DNA, RNA and protein and these mole-cular events are related to the oncogenic process as evidenced by considerable research on the known PAH carcinogens such as benzo[a]pyrene (18). Recent results in our laboratory have shown that 4-chlorobiphenyl, a model PCB isomer (19) and 4-bromobiphenyl (20), a model polybrominated biphenyl (PBB) isomer, are metabolized by rat and rabbit liver microsomes to yield both metabolites and macromolecular adducts (e.g. Fig. 9). Although these substrates are themselves not important as pollutants their metabolic pathway is comparable with their more highly halogenated homologs. These latter substrates are less reactive than the monohalobiphenyls and necessitated the use of the model isomers.

Another important component of mammalian AHH activity is the inducibility of this system by an array of xenobiotics including many aromatic hydrocarbons, chlorinated aromatic (and non-aromatic) hydrocarbons and drugs. The induction process is presumably a response by the target organism to metabolize and detoxify the foreign material. Again AHH induction has been widely studied in terrestrial species and the process has also been confirmed in aquatic animals. Considerable AHH induction was observed using coho salmon (Onchorhynchus kisutch) pretreated with PCB in the diet and with the water soluble component of crude oil (14), PCB induce AHH activ-ity in channel catfish (Ictalurus punctatus) (15) and crude oils also are AHH inducers in brown trout (Salmo trutta) (16). Thus some of the more important aquatic pollutants induce AHH enzyme activity in aquatic animals and most of the AHH induction data has been determined by measuring the increase in the hydroxy-lation of specific aromatic substrates (usually aniline or benzo[a]pyrene).

Since the enzyme system plays a key role in chemical carcinogenesis it is important to be able to evaluate the effects of chemicals not only as inducers of hydroxylase activity (i.e. metabolite formation) but as mediators of macromolecular binding. A possible confusion in using AHH activity as a potential marker of chemical carcino-genesis lies in the method of determining this enzyme activity; namely by measuring the hydroxylation of aniline or benzo[ajpyrene. Thus one uses a metabolic or potential cellular detoxification reaction with a limited number (usually 1 or 2) of substrates to determine AHH activity and its increase by chemical induction. We have proposed an alternative method to assess AHH activity for specific chemical inducers using the following approach and rationale (21).

304 S. Safe, et dl.

rat microsomes

o

pmole/mg 4

protein

(X 1(Γ2)

rabbit microsomes ο

time (min.)

Fig. 9 Microsomal-mediated binding of A, 4-chlorobiphenyl to rat liver microsomes and B, 4-bromobiphenyl to rabbit liver microsomes (the substrates were radiolabelled with tritium).

Metabolism 305

14

13 J

12 J

6

5

4

3

13

12 J

6 J

5

4

3 -I

2

aniline 4-chloro- LMW hydroxylase biphenyl

hydroxylase

HMW cy to chrome P-448

aniline 4-chloro- LMW HMW cytochrome hydroxylase biphenyl P-448

hydroxylase

Fig. 10 Aroclor 1254 induced microsomes and their metabolism of 4-chlorobiphenyl (Top).

Fig. 11 Firemaster BP-6 induced microsomes and their metabolism of 4-chlorobiphenyl (Bottom).

306 S. Safe, et dl.

i) The chemical inducer is administered to the test animal on days 1 and 3 and the animal is sacrificed on day 6.

ii) The hepatic microsomes are isolated and incubated with a radiolabelled sub-strate and cofactors for 30 minutes.

iii) After incubation the amount of radiolabel incorporated into the ether soluble metabolite (M) fraction (detoxification) the alcohol soluble low molecular weight (LMW) conjugates (detoxification) and the insoluble macromolecular-substrate adducts (HMW) fraction (toxification) are determined.

iv) Using the above data a T/D index which is defined as

^ - MTTM X 100° can be determined for each inducer and substrate. Thus it has been proposed

that the T/D index might be more useful in assessing the potential toxicity of an inducer on the AHH system since it takes into account not only the increase in metabolic activity but the corresponding changes in conjugate and macromolecular substrate adducts. An example of the data obtained using this approach in illus-trated in Fig. 10 and 11. The T/D index obtained for the corn oil control experi-ment using rat liver microsomes was 18 whereas values of 48 and 54 were obtained with the rat liver microsomes induced with the commercial PCB (Aroclor 1254) and PBB (fireMaster BP-6). Studies in progress with isomeric halogenated pollutants show a considerable variation in the T/D index with respect to the structure of the inducer with the 3,3f,4,4T-tetrachlorobiphenyl being the only isomer with an index comparable to that of the commercial Aroclor 1254 mixture.

CONCLUSION

Thus the metabolism of pollutants by a variety of animal species results in their conversion to a range of hydroxylated products. This process is accompanied by the formation of an activated arene oxide intermediate which can result in both meta-bolite and metabolite conjugate formation as well as substrate adducts with critical cellular macromolecules (e.g. proteins and nucleic acids). The latter process is associated with chemical carcinogenesis and it is clear that the meta-bolism of pollutants potentially involves both toxification and detoxification with the balance dependent on the substrate, AHH activity, the amount of AHH chemical inducer present and to a host of other environmental and genetic factors.

REFERENCES

(1) A. Kappas and A.P. Alvares, How the liver metabolizes foreign substances, Sei. Amer. 232, 22 (1975).

(2) D. Ryan, A.Y.H. Lu, J. Kwalek, S.B. West and W. Levin, Highly purified cytochrome P-448 and P-450 from rat liver microsomes, Biochem. Biophys. Res. Commun. 64, 1134 (1975).

(3) A.Y.H. Lu, Liver microsomal drug metabolizing enzyme system: functional components and their properties, Fed. Proc. 35, 2460 (1976).

(4) J.E. Chambers and J.D. Yarborough, Xenobiotic transformation systems in fishes, Comp. Biochem. Physiol. 55C, 77 (1976).

(5) S.M. Sieber and R.H. Adamson (1977), The metabolism of xenobiotics in fish, In Drug Metabolism, D.V. Parke and R.L. Smith, Taylor and Francis Ltd., London, 231.

(6) R.H. Adamson, Drug metabolism in marine vertebrates, Fed. Proc. 46, 1047 (1967).

Metabolism 307

(7) 0. Hutzinger, D.M. Nash, S. Safe, A.S.W. de Freitas, R.J. Norstrom, D.J. Wildish and V. Zitko, Polychloroinated biphenyls: metabolic behaviour of pure isomers in pigeons rats and fish, Science 178, 312 (1972).

(8) S. Safe, D. Jones, J. Kohli, L.O. Ruzo, 0. Hutzinger and G. Sundstrom, The metabolism of chlorinated aromatic pollutants by the frog, Can. J_. Zool. 54 1818 (1976).

(9) M.M. Tulp, C. Sundstrom and 0. Hutzinger, The metabolism 4,4T-dichlorobiphe-nyl in rats and frogs, Chemos. 5, 425 (1976).

(10) G. Sundstrom, 0. Hutzinger and S. Safe, The metabolism of chlorobiphenyls, Chemos. 5, 267 (1976).

(11) M.J. Melancon and J.J. Lech, Isolation and identification of a polar metabo-lite of tetrachlorobiphenyl from bile of rainbow trout exposed to 1I+C-tetrachlorobiphenyl, Bull, Environ. Contam. Toxicol. 15, 181 (1976).

(12) D.M. Jerina and J.W. Daly, Arene oxides: a new aspect of drug metabolism, Science 185, 573 (1974).

(13) J.T. Ahokas, 0. Pelkonen and N.T. Karki, Metabolism of polycyclic hydro-carbons by a highly active aryl hydrocarbon hydroxylase system in the liver of a trout species, Biochem. Biophys. Res. Commun. 63, 635 (1975).

(14) E.H. Gruger Jr., M.M. Wekell, P.T. Numoto and D.R. Craddock, Induction of hepatic aryl hydrocarbon hydroxylase in salmon exposed to petroleum in seawater and to petroleum and polychlorinated biphenyls, separate and together in food, Bull. Environ. Contam. Toxicol. 17, 512 (1977).

(15) D.W. Hill, E. Hejtmancik and B.J. Cump, Induction of hepatic microsomal enzymes by Aroclor 1254 in Ictalurus punctatus (channel catfish), Bull. Environ. Contam. Toxicol. 16, 495 (1976).

(16) J.F. Payne and W.R. Penrose, Induction of oryl hydrocarbon (Benzo[a]pyrene) hydroxylase in fish by petroleum, Bull. Environ. Contam. Toxicol. 14, 112 (1975).

(17) C. Heidelberger, Chemical carcinogenesis, Ann. Rev. Biochem. 44, 79 (1975).

(18) R.I. Freudenthal and P.W. Jones, Eds. (1976), Polynuclear Aromatic Hydro-carbons : Chemistry, Metabolism and Carcinogenesis, Raven Press, N.Y.

(19) C. Wyndham and S. Safe, The In vitro metabolism of 4-chlorobiphenyl by control and induced rat liver microsomes, Biochem. in press.

(20) J. Kohli, C. Wyndham, M. Smylie and S. Safe, The metabolism of bromobiphenyls, Biochem. Pharmacol., in press.

(21) A. Crawford and S. Safe, An assessment of the effects of enzyme inducers on aryl hydrocarbon hydroxylase activity, Res. Commun. Chem. Pathol. Pharmacol., in press.

Laboratory Microcosms for Use in Determining Pollutant Stress

FRANK G. WILKES

Environmental Protection Agency, Environmental Research Laboratory Gulf Breeze, Florida 32561

INTRODUCTION

The regulatory actions of the Environmental Protection Agency must be based on sound scientific evidence. Historically, the regulation of the introduction of pollutants into the aquatic environment has relied on data from acute bioassay tests using a single species. Results from chronic (sublethal) tests have been utilized although the number of aquatic species available for such tests are relatively low in number, particularly in the marine environment. Both acute and chronic bioassays emphasize the direct effects of pollutants on various life stages of aquatic organisms. In order to more accurately describe the effects of pollutants in aquatic systems, tests must be developed to assess indirect or ecological effects of pollutants. Such tests would determine pollutant effects on population, community, or ecosystem level parameters as well as effects on the interactions among organisms and their physical/chemical surroundings. Data from such tests would not only be valuable in describing sublethal estuarine pollutant effects, but would also aid in placing single species toxicity data in the per-spective of real-world ecosystems dynamics.

Development of such tests for investigating estuarine ecosystem compartments and processes is a major objective of the Environmental Research Laboratory, Gulf Breeze, Florida. An arsenal of tests suitable for investigating different com-pounds and generating information applicable to unique problems is being produced by utilizing a multidisciplinary approach. Potential uses of the data from these methods include:

. revision and establishment of water quality criteria;

. revision and establishment of pesticide registration and application guide-lines;

. evaluation of potential substitute chemicals;

. establishment of toxic substance protocols;

. establishment of ocean dumping guidelines;

. establishment of dredged material disposal guidelines.

The suitability of these tests for the above applications is of paramount impor-tance in their development. Further, factors must be established to determine how a test or combination of tests should be selected.

Contribution No. 357, U.S. Environmental Protection Agency, Gulf Breeze, Florida

309

310 F. G. Wilkes

Implicit in the concept of developing tests for individual ecosystem compartments and processes is the realization that one test alone will not yield definitive information on the fate and effects of a pollutant in a system as complex as an estuarine ecosystem. Consequently, an attempt must be made to define the compart-ments and processes of estuarine systems and to develop tests applicable to each. Fig. 1 depicts a number of components and processes which occur in the water

Fig. 1. A diagramatic representation of a number of broad ecosystem processes and components.

column and at air-water and water-sediment interfaces, meant to be representative rather than inclusive.

The processes shown are

As shown on Fig. 1, the estuarine ecosystem is composed of three media--air, water and sediment. Important processes occurring within each media and their mutual boundaries must be considered. The biological components of the ecosystem, represented by different trophic levels in Fig. 1, occur primarily in the water column. Important microbiological processes are found at air-water and water-sediment interfaces. Physical and chemical processes throughout the system deter-mine the distribution and form of pollutant to which organisms are exposed and mediate and control many of the biological processes. Such physical processes include volatilization, solubilization and sorbtion. Methods for determining the extent and significance of these processes are available for most compounds.

Upon introduction into estuarine ecosystems, a xenobiotic, or foreign compound, may interact in a number of ways. For example, it may be accumulated directly by organisms of all trophic levels. This is represented by the left vertical column on Fig. 1. Direct accumulation represents the accumulation by each organism from all sources and may be broken down into accumulation directly from food and di-rectly from water. Methods to determine the accumulation by estuarine organisms of pollutants via various routes have been reported by numerous investigators

Laboratory Microcosms 311

(1, 2, 2). Direct accumulation also occurs with lower trophic level organisms, as shown in Fig. 1,

In addition to direct accumulation by individual organisms, toxicants may be bio-accumulated in food chains or through food webs. This process produces in upper trophic leyel organisms higher chemical concentrations than would be accumulated through direct uptake alone. This is represented on Fig. 1 by the compartment "Food Chain Bioaccumulation." Methods of determining the bioaccumulation of organic chemicals through estuarine food chains have been developed by Bahner et al. 0 ) .

The middle yertical column in Fig. 1 depicts the direct effects of pollutants on organisms of different trophic levels. Direct effects refer to the direct action of pollutants on individual organisms and include such effects as mortality, growth and reproductive effects, physiological and pathological effects, and behavioral effects. Data from bioassay tests, for which mortality is the response of inter-est, have long aided the establishment of water quality criteria and other regula-tory guidelines (4, 5, 6}. Likewise, tests designed to dete'^ine the effects on aquatic organisms of long exposures to sublethal concentrations of pollutants have been useful to regulatory agencies (7, 8, 9). Less obvious physiological and pathological damage to organisms exposed to low toxicant levels has been observed and tests have been developed to quantify these effects (10, 11, 12).

Relatively fewer studies have been conducted on the effects of pollutants on behav-ior. Such information is important because behavioral responses usually offer a much more sensitive indicator of pollutant effects than mortality. Behavioral effects may be indicative of subtle changes in important ecosystem interactions and thus indicative of responses by multiple rather than single species. One such behavioral test, which demonstrates the effect of the pollutant methyl parathion on the feeding activity of the lugworm Arenicola cristata, is described herein.

The right vertical column in Fig. 1 depicts secondary, or indirect, effects of pollutants on various trophic levels. Indirect effects are exerted by sublethal concentrations of pollutants on the processes linking organisms within the same community. Such effects, whether or not ultimately lethal to some organism groups, are certain to produce a change in the ecology of the system. Whether this change is significant, or even harmful, depends on the significance of the process af-fected and the magnitude of the stress applied to the system. A test developed to investigate the effects of methyl-parathion on an environmentally important predator-prey relationship is described herein.

As shown on Fig. 1, microbial processes take place at air-water and water-sediment interfaces as well as in the sediment itself. At both of these sites of high microbiological activity, pollutants introduced into the estuarine environment may affect the metabolism of the microorganisms themselves, or, in turn, may be broken down by the microorganisms into various degradative products. A microbiological microcosm containing both water and sediment has been developed to investigate these interactions.

The term ecosystem response, as used on Fig. 1, refers to total ecosystem func-tional and structural changes caused by the introduction of a pollutant and is determined by measuring productivity, species diversity, population dynamics, etc. A microcosm designed to investigate the dynamics of salt-marsh ecosystems is described herein.

312 F. G. Wilkes

MICROCOSMS

Benthic System

One of the major concerns of regulatory agencies is the determination and predic-tion of the effect of pollutants on benthic organisms. Such information is neede< to determine both the potential impact of chemicals, which move from the water column into sediments, and of dredged material deposited on the bottom.

Marine polychaetes produce distinct, characteristic features on a substrate sur-face as a result of their activity. For example, the lugworm Arenicola cristata produces funnel-shaped depressions on the substrate surface as a result of its normal feeding processes. A decrease in the number of feeding funnels produced b) the lugworms would indicate an interruption in this activity. Experiments have been conducted at the University of West Florida to develop a method of monitorin« changes in these surface features as a function of environmental stress (13).

Tests were conducted in two static, aerated, 125-liter aquaria (Fig. 2). Each

Fig. 2. Diagram of Benthic System (A. 24 hr timer; B. 35 mm camera; C. lights; D. and E. 125 I aquaria with 25 cm sand and 75 I seawater). From Rubinstein (13).

aquarium received 25 cm of clean sand and 72 liters of filtered seawater. All tests were conducted at salinities between 20 and 25 /oo and temperatures betweer 18 and 22°C.

Lugworms of similar size were placed in each tank and 70 gr of ground seagrass wer then added, forming a dark mat approximately 3 mm thick over the sediment surface.


After the system had acclimated, the test compound, methyl parathion, was intro-duced into one tank. The non-dosed tank served as a control.

Changes in the surface features produced by the lugworm were monitored by time-lapse photography. A photograph of the substrate surface was taken at 12-hour intervals for 144 hours. The amount of surface disturbance was determined at each 12-hour interval. The cumulated surface area disturbed was plotted against time to provide a substrate modification rate for both exposed and control lugworms. Since both tanks were treated identically, any significant difference in the rate of substrate modification could be attributed to the test compound.

The results of exposure of the lugworms to three concentrations of methyl parathion are presented in Fig. 3. At 25 and 75 yg/£ there was no difference between the

3O0r

25 i ig/l METHYL-RARATHION

200 [ ·

l00| . ^^*^i*-''''"*>

0« ■ ■ ■ ■ ■ ■

300

200

*, 100

§ 400

i 300

200

100

° 0 84 96 ΙΟβ 120 132 144

TIME (hour«)

Fig. 3. Comparison of rates of sediment turn-over by exposed (X) and control (C) lugworms at three methyl parathion concen-trations. From Rubinstein (13).

substrate disturbance rate of the exposed and the control lugworms. At 150 ml the feeding activity of the exposed lugworms was significantly lower than that of the control.

Based on these results, a water concentration between 75 and 150 yg/£, or lower, would interrupt the normal cyclical pattern of the lugworm resulting in a decrease in its substrate reworking activity. This could reduce the lugworm-mediated exchange between the water and sediment resulting in an increased residence time in the water of the pollutant. Reduced feeding activity of the lugworm could also ultimately result in the death of these organisms. Such an event would affect the overall transport of nutrients and pollutants throughout the system, as well as alter food chains in which the lugworm forms a part.

314 F. G. Wilkes

Whatever the environmental significance of this deviation, it is clear that this technique can demonstrate a behavioral effect on an important benthic organism at sublethal concentrations. Such a test, which demonstrates the response of an organism to low levels of pollutant, will be of value in providing a yery sensi-tive means of determining the potential impact of substances introduced into the estuarine environment on an important ecological process.

Predator/Prey System

Estuarine ecosystems contain many important species interactions that are vital to the structural and functional integrity of the system. Many of these processes may be vulnerable to stress and may be altered significantly by pollutants, thus causing the ecosystem to deviate from its stable state. Because they play a major role in determining community structure and species diversity, many predator/prey relationships fall in this category of vulnerable ecosystem processes. A test or method that would demonstrate and quantify the effect of a pollutant on an impor-tant estuarine predator/prey relationship would therefore be valuable in predicting the potential impact of a contaminant on an estuarine ecosystem. Such a test, de-signed to determine the effect of an organic pollutant on a specific estuarine predator/prey relationship, has been developed at the University of West Florida (14).

It has been demonstrated that methyl parathion impairs the ability of grass shrimp Palaemonetes pugio to escape predation by the Gulf killifish Fundulus grandis (15). It is also known that the juvenile sheepshead minnow Cyprinodon variegatus, another major prey of £. grandis, is much more tolerant of organophosphate pesticides, based on 96-hr LC50 values, than are grass shrimp (16, 17). In a multi-prey sys-tem, the predator £. grandis may be expected to consume a higher proportion of affected or impaired species in the presence of an organophosphate pesticide. An experiment was conducted, therefore, to determine if low concentrations of methyl parathion might affect the rate of predation and prey preference of £. grandis when provided with two prey, £. pugio and £. variegatus, simultaneously.

After acclimation in the laboratory for 14 days, 24 adult £. pugio and 24 juvenile £.· variegatus were exposed to three concentrations of methyl parathion in 160-liter, aerated aquaria. Following 24-hr exposure, a single IF. grandis was intro-duced into each tank. The number of individuals of each prey species remaining in the aquaria were counted each day for five days following introduction of the predator. The ratio of £. pugio to £. variegatus surviving after each day of predation at each toxicant concentration was calculated.

The results of this experiment are shown in Fig. 4. The ratio of shrimp to fish in the control aquaria increased rapidly during the five-day test, indicating a strong preference of IF. grandis for _C. variegatus. At a concentration of 0.021 vq/l, the ratio of shrimp to fish increased until day 4 after which it leveled off, indicating that the predator had switched from fish to shrimp. A similar switch occurred at 0.119 yg/£ and at 0.475 yg/£ on days 3 and 2, respectively, indicating that the higher the methyl parathion concentration, the more rapid the deliterious effect on the shrimp's escape behavior. After five days, the methyl parathion effect on predation was clearly concentration dependent: increased methyl parathion concentration resulted in increased preference of £. grandis for £. pugio relative to £. variegatus.

These results indicate that sublethal concentrations of methyl parathion can alter the relative proportions of prey in a predator's diet. Given a choice of £. pugio and juvenile £. variegatus, both exposed to methyl parathion, adult £. grandis increased the proportion of £. pugio in their diet over time. This change in predator preference became more pronounced with increasing concentration of

Laboratory Microcosms 315 methyl parathion, and was apparently caused by a decrease in the ability of P_. pugio to avoid predation.

x 3 CO

CONTROL /

02\ jjg/l

\ QMS jigΛ

Fig. 4. Ratio of two prey species remaining after predation by Fundulus grandis in different concentrations of methyl parathion. From Farr (14).

Although it is difficult to extrapolate results obtained in this laboratory experi-ment to what might take place in nature, these results suggest that low levels of methyl parathion in an estuary could result in a reduction in number of individual crustaceans and an increase in number of smaller fishes, including juveniles. The increase in fish density and the resultant increase in competition could alter growth and survival patterns of ecologically and commercially important fish species. Such a long-term detrimental impact cannot be predicted from the results of this test. The method can be .used, however, to screen the effects of chemicals on this particular ecosystem process. It will provide information which, although not of a magnitude and significance necessary to form the complete basis for deci-sions on the environmental compatability of potential pollutants, may aid when added to the results of other studies in the total assessment of the ecological damage which might be expected when certain chemicals are introduced into the estuarine environment.

Eco-Core Microcosm

The Eco-Core, an artificial ecosystem or microcosm, has been developed at the Environmental Research Laboratory, Gulf Breeze as a method for determining the fate and effects of organic pollutants particularly as they relate to the activi-ties of microorganisms under conditions which simulate the natural environment (18). The Eco-Core consists of a glass tube, which can be used to core a sample of water and sediment and then returned to the laboratory (Fig. 5). The pollutant

316 F. G. Wilkes

AIR OR „ N I T R O G E N ^ C 0 2 ABSORBENT

' COLUMN (Optional)

ORGANIC RESIN TRAP

MANIFOLD FOR MULTIPLE

CORES

Ϊ Γ Ε Χ Ι Τ

_ j £ i y ^ L _ _

T SEDIMENT

-ALKALI TRAP

-CONSTANT TEMR ENVIRONMENT

SILICON STOPPER

Fig. 5. Eco-Core Microcosm (18)

is then introduced and the system aerated. The fate of the parent compound can be followed, the accumulation of degradation products can be monitored, and the effect of the pollutant contained therein on the bacterial populations can also be deter-mined. This type of Eco-Core System permits a large number of replicates to be employed and the effects of a large number of environmental variables tested on the degradation process.

Radiolabeled pollutants are used to facilitate analysis. Water column degradation products are continuously monitored by-.thin-layer chromatography and autoradio-graphy. Volatile products, including CCL and organics, are continuously scrubbed from air exiting the systems. The effects of the toxicant on the microbial ecolo-gy of the cores are monitored by heterotrophic plate counts and total COo evolu-tion. Changes in microbial physiological indices induced by pollutants are ob-tained by selective media plating from the water column.

From these studies a number of important considerations about the microbial degra-dation of a pollutant have evolved. First, a total budget analysis approach is essential...In the case of methyl parathion, for example, determining just the amount of CCL released would be yery misleading as the CCL produced represented only 5% of the original methyl parathion concentration, thus indicating relatively little degradation. Examination of other degradation products, however, revealed that extensive degradation (90%) of the methyl parathion had occurred.

Second, these studies reveal the critical nature of incorporating all major com-ponents of the natural aquatic environment into a laboratory system. Again when using methyl parathion, sediment plays a very critical role in initiating the deg-radation process. In Eco-Cores containing only water (no sediment), degradation of methyl parathion was considerably slower than with sediment, and the extent of degradation was not as great. However, the results indicate that once the degra-dation process is initiated in the sediment, further degradation occurs in the water column. Thus, the water compartment of the ecosystem is equally important.


Third, the simplicity of a model ecosystem like the Eco-Core offers a number of advantages which make a study on the Eriodegradation of a xenobiotic compound con-siderably more valid. With methyl parathion, it has been shown that this replica-bility of the systems, the minimal effort involved in setting up large numbers of cores, the ease of handling and analyzing the systems and the use of these systems over long incubation periods have contributed intensely to understanding the fate of methyl parathion in aquatic salt-marsh environments.

Limitations to these microcosms because of their static, closed nature and the possible antifactual changes which might be induced by accumulating metabolites are recognized, however. Actual reaction rates will have to be established in systems that more closely model the dynamic nature of an aquatic system. To solve these problems, systems which incorporate continuous flow characteristics are being developed as a logical extension in microbiological microcosm technology.

Salt-Marsh Ecosystems

At the University of South Carolina Belle Baruch Laboratory, a study is underway to develop and test replicate experimental salt-marsh units at the microecosystem lev-el as diagnostic tools for the assessment of both long and short-term pollution effects on the Spartina salt-marsh community (19). Data from these units will be compared with data obtained from natural salt marshes and will be utilized in the development of mathematical models of salt-marsh ecosystems.

The system consists of 4 square tanks (8'x8"), program timers, pumps, solenoid valves, and pipe plumbing (constructed of PVC plastics or noncorrosive metals) which provide the plots with daily tides. The source waters are from a natural creek adjacent to the site of marsh plot extractions which flow into and out of the tanks at rates which simulate with temporal synchrony the height and period of inundation of the site from which the plots were extracted.

The community metabolism response was selected as the primary test response. The reasons for selecting this response are: 1) It has been shown to be a yery sensitive indicator of community imbalances due to stress in aquatic ecosystems; 2) Replicate laboratory aquatic systems do not differ significantly with respect to levels of community metabolism, and 3) The automation of response measurement is facilitated.

The findings of the microcosms studies are also being projected to the natural sys-tem. This is vital to court deliberations relating microecosystem results to the macrosystem. To accomplish this: 1) Studies have been initiated of both the micro and macro infauna community structures in order to compare and contrast the test plots with the natural system; 2] The plots were extracted with substrate intact to a depth of 18 cm to insure incorporating entire faunal and floral commu-nities, and 3) Natural species and nutrient recruitment is accomplished through the tidal systems.

The microecosystem community metabolism studies have shown very good agreement among replicates. A nutrient flux determination technique has been developed. The ability of the marsh plots to either assimilate or export nitrogen and phos-phorous nutrients to the waters coursing in and out of the marshes is measured directly. Preliminary results indicate the marshes are "sinks" for the "sources" of N. These studies are the first direct measurement of fluxes of this sort for marshes.

These microecosystems incorporate several inherent advantages because they closely approach the natural system. Among the advantages of this system are:

318 F. G. Wilkes

A] The systems were extracted plots with intact substrata and fauna! assemblages. BJ The ttdal seawater system utilizes a natural tidal creek adjacent to the plot extraction site, as a source of water for the flowthrough system. This encourages natural species and nutrient recruitment. The tides are in synchrony temporarily with the natural site, and simulate the period and height of inundation of the natural site.. The enclosures and plumbing allow for ease of sampling and flux de-terminations. Cl The iDicroecosystems are open to the atmosphere and are subject to natural re-gimes of light, temperature and rainfall common to the natural marsh. D\_ The plot size, 6m2, minimizes any "edge" effect and ensures for suitable re-sponse measurements minimizing scaling errors. E) Plots can he sacrificed and replaced with relative ease. F) No idigenous greenhouse insect pests are present. Insect populations closely resemble the populations of the natural marsh site.

SUMMARY

The goal of this research is to develop a number of tests that will yield informa-tion about different processes and mechanisms within estuarine ecosystems: tests that will examine the non-biological and the biological, the fate and the effects. These tests may be termed "microcosms," as microcosms have been defined as minia-ture ecosystems containing components and processes necessary to investigate spe-cific origins, flows, fates and/or effects of materials in the environment (20). Microcosms are used to represent segments of the environment and to integrate those interactions among ecosystem processes and components not obtainable in single species experiments. Tests described in this discussion focus on particu-lar processes and particular components of the estuarine environment, i.e., par-ticular sub-units of the estuarine ecosystem. All tests are not applicable to all compounds. By selectively choosing tests on the basis of the compound in question and the information desired, it will be possible to develop information applicable to a particular compound and the problem at hand.

Fig. 6 depicts an array of such tests. The tests vary in complexity. They also vary as to the different points, components, and processes in the ecosystem in which they focus. Therefore, depending on the information needed, any number or combination of these tests may be selected to form a protocol designed to provide the required information. Further relationships among these tests have been previously described (21).

The objective in developing these techniques is to provide the Agency with methods to obtain information necessary for its regulatory activities. The tests them-selves may be used as part of a screening process or protocol for the identifica-tion of potentially hazardous compounds. Data from these tests will be used in the establishment of water quality criteria, toxic substance protocols, pesti-cide registration guidelines, effluent guidelines, and ocean-dumping guidelines.


XENOBIOTIC

i Transformation

NONBIOLOGICAL I

BIOLOGICAL

1 Compartmentatization

Photolysis Hydrolysis Chelation Volati l i ty Adsorption Solubility

I Macrobiological

Lab Microcosm

HBenthic

I— Marshland

h-Estuarine

L-Gulf 1

Microbiological I

Fate I

Effects

Metabolism Accumulation

Physiological

Pathological I

Biochemical

I Behavioral

, I ,

Fate

_L_ Effects

I I I Carbon I Metabolism Accumulation Diversity dioxide Population

I I I Predator-Activity Avoidance Preference prey

Community I

— i Toxici ty

Diversity Succession Acute Chronic

Fig. 6. Tests to determine the effects of xenobiotics on ecosystem processes and components.

ACKNOWLEDGMENTS

The Predator/Prey System and the Benthic System were developed by J. A. Farr and N. I. Rubinstein, respectively, under an ERL, GB grant to the University of West Florida, C. N. D'Asaro, Principal Investigator. The Salt-Marsh System was devel-oped by W. Kitchens under an ERL, GB grant to the University of South Carolina, F. J. Vernberg, Principal Investigator. The Eco-Core microcosm was developed by A. W. Bourquin, M. A. Hood, R. L. Garnas and P. H. Pritchard, Environmental Research Laboratory, Gulf Breeze.

REFERENCES

1. L. H. Bahner, J. M. Sheppard, J. M. Patrick, L. R. Goodman and G. E. Walsh, Kepone bioconcentration, accumulation, loss and transfer through estuarine food chains, Chesapeake Sei. 18, 299 (1977).

2. D. J. Hansen, P. R. Parrish and J. Forester, Aroclor 1016: Toxicity to and uptake by estuarine animals, Environ. Res. 7, 363 (1974).

3. S. C. Schimmel, J. M. Patrick, Jr. and J. Forester, Heptachlor: Uptake, depuration, retention and metabolism by spot, Leiostomus xanthurus, J. Toxicol. Environ. Health 2, 169 (1976).

4. J. I. Lowe, Effects of prolonged exposure to Sevin on an estuarine fish, Bull. Environ. Cont. Toxicol. 2, 147 (1967).

5. D. R. Nirmio and L. H. Bahner (1976). Metals, pesticides and PCB'siToxicities to shrimp singly and in combination. Irv. Estuarine Processes Vol. I. Uses, Stresses and Adaptation to the Estuary, Academic Press, New York.

320 F. G. Wilkes

6. W. A. Brungs, J. R. Geckler and M. Gast, Acute and chronic toxicity of copper to the fathead minnow in a surface water of variable.quality, Water Res. 20, 37 (1976).

7. J. I. Lowe, P. D. Wilson, A. J. Rick and A. J. Wilson, Jr., Chronic exposure of oysters to DDT, toxaphene and parathion, Proc. Nat. Shellfish Assoc. 61, 71 (1971).

8. D. J. Hansen, S. C. Schimmel and J. Forester, Aroclor 1254 in eggs of sheeps-head minnows: effect on fertilization success and survival of embryos and fry, Proc. 27th Annu. Conf. Southeast Assoc. Game Fish Comm. (1973).

9. J. G. Eaton, Chronic cadmium toxicity to the bluegill (Lepomis macrochirus Rafinesque), Trans. Am. Fish. Soc. 103, 729 (1974).

10. J. A. Couch and D. R. Nimmo, Ultrastructural studies by shrimp exposed to the pollutant chemical polychlorinated biphenyl (Aroclor 1254), Bull. Soc. Pharmacol. Environ. Path. 11, 17 (1974).

11. D. L. Coppage and E. Matthews, Brain-acetylcholinesterase inhibition in a marine teleost during lethal and sublethal exposures to 1,2-dibromo-2, 2-dichloroethyl dimethyl phosphate (naled) in seawater. Toxicol. Appl. Pharmacol. 31, 128 (1975).

12. R. A. Drummond, G. F. Olson and A. R. Batterman, Cough response and uptake of mercury by brook trout, Salvelinus fontinalis, exposed to mercuric com-pounds at different hydrogen-ion concentrations. Trans. Am. Fish. Soc. 103, 244 0974).

13. N. I. Rubinstein, A benthic bioassay utilizing time-lapse photography to measure the effects of toxicants on the feeding behavior of lugworms (polychaeta:Arenicolidael. Proc. Sympos. Pollut. and Physiol. Mar. Organisms (In press).

14. J. A. Farr, The effect of methyl parathion on predator preference for two estuarine prey species, Trans. Am. Fish. Soc. (In press).

15. J. A. Farr, Impairment of antipredator behavior in Palaemonetes pugio by exposure to sublethal doses of parathion, Trans. Am. Fish. Soc. (In press).

16. R. Eisler, Acute toxicities of insecticides to marine decapod crustaceans, Crustaceana 16, 3Q2 096?).

17. R. Eisler, Acute toxicities of organochlorine and organophosphorus insecti-cides to estuarine fishes, U.S. Fish. Wildl. Serv. Tech. Paper 46, 1 (1970).

18. A. W. Bourquin, M. A. Hood and R. L. Garnas (1977). An artificial microbial ecosystem for determining effects and fate of toxicants in a salt-marsh environment, In Development in Industrial Microbiology Vol. 18.

19. F. J. Vernberg, R. Bonnell, B. Coull, R. Dame, Jr., P. DeCoursey, W. Kitchens, Jr., B. Kjerfve, H. Stevenson, W. Vernberg and R. Zingmark, The dynamics of an estuary as a natural ecosystem, U.S. Environ. Prot. Agency, Ecol. Res. Ser., EPA-600/3-77-016 (1977).

20. National Science Foundation, Ecosystem processes and organic contaminants: Research needs and an interdisciplinary perspective (n.d.).


21. A. W. Bourquin, R. L. Garnas, P. H. Pritchard, F. G. Wilkes, C. R. Cripe and N. I. Rubinstein, Interdependent microcosms for the assessment of pollu-tants in the marine environment, Int. J. Environ. Studies (In press).

European Regulatory Actions Relating to Aquatic Pollutants

D. J. DE GEER Division of International Affairs, Ministry of Health and Environmental Protection, Leidschendam, Holland

"What of thee I dig out, let that quickly grow over,

Let me not hit thy vitals, or thy heart"

(Hymn to Earth from the Atharva Veda)

Those striving to preserve the quality of the water in Europe need to know the

European regulations. In general individual countries in Europe are too small to

solve major environmental problems themselves. The Rhine, to give you a well-

known example, courses through four different countries. To give you another in

the field of air pollution: one of the main conclusions of a recent OECD report

on air pollutants is that sulphur compounds travel long distances in the atmosphere

(several hundred kilometres or more) and that the air quality in any one European

country is measurably affected by emissions elsewhere in Europe. This does not

mean that we have to wait for international regulations to start. On the contrary,

also what we regard as "European problems" in the first instance have to be ap-

proached nationally. Another way of approaching environmental problems at European

level is through European integration. Government leaders at a European summit in

October 1972 in Paris decided that there should be a European environmental policy.

I quote:

"Economic expansion is not an end in itself. As befits the genius of Europe

particular attention must be given to intangible values and to protecting the en-

vironment, so that progress may really be put at the service of mankind."

I shall come back to the Rhine and to the Common Market.

323

324 D. J. de Geer

Before doing so I should like to place my remarks in a historical context and I

also think it would be useful to go into the reasons for Holland's interest in

environmental questions. After looking at what has already happened in connection

with the Rhine and the Common Market, I shall finally say a few words about the

developments we expect and what our priorities must be.

People were interested in water long before 1972.

The Greek philosopher Thales of Miletus considered water to be the origin of

creation and some centuries later Emperor Ashoka of India drafted environmental

legislation containing a number of regulations of which we could be envious.

Before jumping to "the international environment year" of 1972, I mention the

Rhine salmon convention of 1885 which clearly did not meet its purpose and the

Council of Europe.

The Council of Europe recognized the seriousness of water pollution at a very

early stage. A Charter on water was made in May 1968 establishing a number of

principles which 10 years later are just as valid.

It emphasized that besides maintaining an acceptable quality of water depending

on the different uses of a stream, special attention must be paid to safeguarding

the natural properties of the water. At this stage the ecological approach was

already recognized.

A second point which in practice is turning out to be highly significant was that

water quality has to be dealt with on the scale of the natural water basins.

In 1972 there was the Stockholm conference in which the developing countries too

accepted that environmental protection is important. The former prime minister of

India, Mrs. Indira Gandhi, made a memorable speech in which she told about the

historical unity of man and environment in her country and said that environ-

mental pollution is a cultural problem.

I quote: "Pollution is not a technical problem. The fault lies not in science and

technology as such but in the sense of values of the contemporary world which

ignores the rights of others and is oblivious of the longer perspective".

She also warns that developing countries have and must have a different

appreciation of the relation between environment and economic development.

I quote: "The rich countries may look upon developments as the cause of environ-

mental destruction, but to us it is one of the primary means of improving the en-

vironment for living, of providing food, water sanitation and shelter, of making

the deserts green and the mountains habitable. We see that however much man

European Regulatory Actions 325

hankers after material goods, they can never give him full satisfaction. Thus the

higher standard of living must be achieved without alienating people from their

heritage and without despoiling nature of its beauty, freshness and purity so

essential to our lives".

I myself am convinced that an effective world-wide environmental policy cannot

exist before a fairer distribution of material wealth has been achieved.

Also in 1972 the World Health Organization made an important contribution with the

publication of a reference work on the dangers to health arising from environ-

mental pollution. Criteria were worked out.

On the European level the Rhine pollution was accepted as a major political pro-

blem by the first Ministerial Conference on the Rhine in The Hague 1972. And the

Oslo convention for the prevention of marine pollution by dumping in the North Sea

area was signed. It was the first time a distinction was made between the various

pollutants, which were divided into three categories.

I am often asked to explain the great interest in environmental questions in the

Netherlands.

A simple answer is there are more than 500 environmental action groups and or-

ganizations keeping the government on their toes. Others seek an explanation in

the psychological characteristics of the Dutch, either their tendency to ponti-

ficate or the attentiveness needed to keep our country a fit place to live in.

I have tried to find some more criteria which explain the degree of active in-

terest. I picked on population density, industrial production per square kilo-

metre and the number of private cars per square kilometre. These may give an in-

dication of the strain which man and his machine place on the environment.

The Netherlands has 370 inhabitants per square kilometre as oppose to 230 in the

united Kingdom and 23 in the U.S.A.

Our industrial production of 300,000 dollars per square kilometre is only exceeded

by Germany's (the United Kingdom's is half of this and the U.S.A.'s just over 101

of it). With 88 private cars per square kilometre we are at the top of the list

(the U.S.A. has only 10 per square kilometre).

Furthermore we are forced to look across our frontiers because our country is al-

most entirely a product of the three major international rivers the Rhine, the

Maas and the Scheldt. Almost 751 of the water available in the Netherlands origin-

ates abroad, of which the Rhine alone accounts for 90%.

326 D. J. de Geer

I promised to tell you something about the developments concerning the Rhine since

the ministerial conference of 1972. I shall not burden you too much with figures

on the Rhine basin. We have put some data down on paper to give you an impression

of the size of the problems at the moment and of the problems which we will en-

counter in the future. The majority of these data have been derived from the long

term work programme which gives an analysis of the present position and puts for-

ward plans and hypotheses up till the year 1985 or in some cases the year 2000.

One wonders whether the situation facing those who have to conduct an environ-

mental policy for the Rhine basin, precludes a fundamentally satisfactory solution.

It might be that there are too many people (30 million) living and working in the

drainage basin of the Rhine to maintain the natural ecological balance properly

along with our current high standard of living.

Why was it necessary in 1972 to talk about Rhine pollution at ministerial level?

The international commission for the protection of the Rhine against pollution,

set up by the 1963 Berne Agreement, by virtue of its terms of reference had been

unable to reduce the salinity of the Rhine and the general quality of Rhine water

was deteriorating; 20 million tons of waste are carried down inside the river

each year (compared with transportation of 120 million tons of useful cargo over

the Rhine).

The 1972 Conference accepted a general plan to restrict the salinity of the Rhine.

The Potassium mines in the Alsace were to limit their saline discharges by

60 kilogrammes per second, i.e. about half. In December 1976 a convention was

signed in which these agreements were laid down. The reduction of 60 kilogrammes

per second has to be achieved in three stages before 1980. The cost is to be borne

jointly by the partners.

An important agreement at the conference was that future power stations will be

equipped with closed cooling systems or other systems with similar effects. The

significance of this decision may be found in the survey of plans for expanding

energy production along the Rhine river. A convention on thermal pollution will

be drawn up.

Finally in 1972 the first step was taken towards what would become the Convention

for the Protection of the Rhine against Chemical Pollution. Agreement on this con-

vention was reached in April 1976 and the convention was signed in December.

I have had the opportunity to be closely involved in the negotiations of the

so-called "chemical convention". And I would like to tell you about it. During

the first year lists of substances similar to those for the Oslo Convention and

a phased action programme were drawn up fairly quickly. Was it because technical


experts were doing the job and "negotiating people" did not come in yet? Is not it

a tempting conclusion? But why, in fact, negotiations became more difficult? My

impression is that the economic recession resulting from the oil crisis reduced

willingness to take on far-reaching environmental commitments. A second important

factor was the fear of primarily the chemical industry in the Rhine area, that the

cost of a quick and effective anti-pollution policy for the Rhine basin would make

them less competitive with other locations (they forget about the enormous ad-

vantages they have of being located in the Rhine basin area). Negotiations went

through a tentative and vague period which continued until October 1974 when the

European Commission made proposals for a community policy to implement the various

international conventions which had been signed or which were still to be con-

cluded. The proposals were structurally the same as the drafts for the Rhine Con-

vention.

It was becoming increasingly clear that for Germany a European regulation was a

prerequisite for willingness to conclude a special convention on the Rhine.

A series of discussions and negotiations at top level during 1975 and at the be-

ginning of 1976 finally led to agreement, which can be summarized as follows:

The Communityfs policy is to provide a framework and must help to ease effects on

competition; policy on the Rhine will concentrate on cleaning up the area.

The Chemical Convention contains two categories of substances, a black list of the

most dangerous and a grey list of less dangerous substances. You will find both

lists in the background material. The aim of the convention is to end pollution

of the Rhine by substances of the black list and reduce pollution by substances

of the grey list.

The uses to which the Rhine is put are laid down in the convention. These are

non-quantified objectives which are important touchstones for the measures to be

taken and also form a basis for establishing quality objectives, that means

quantified objectives for the quality of the surface water.

The first requirement is that the Rhine has to be suitable for processing as drink-

ing water for human consumption. In the EEC - and I shall return to this shortly -

standards for surface water to be used for drinking have been established. This

means that the standards and quality objectives which apply in the EEC are going

to apply automatically to the Rhine. The other requirements can be found in your

documentation.

Rhine policy furthermore has deliberately been placed in the context of the pro-

tection of the seas and oceans. This objective finds its logical conclusion in the

agreement that Rhine anti-pollution policy will stretch to the coastline. Rotter-

dam and its industrial area therefore comes under the regulations of the Chemical

328 D. J. de Geer

Convention.

The measures in the convention to stop or restrict pollution of the Rhine by sub-

stances of the black or grey list, relate in the first place to the disposal of

industrial and domestic waste water.

A question which played a major role in the negotiations was whether we should

concentrate on attaining a certain water quality or only on measures against dis-

charge of waste.

The Netherlands has always maintained that quality standards for Rhine water have

priority and that the effect of measures against discharges must be judged in this

light.

Other place more emphasis on uniform measures against discharges of this kind.

They are opposed to a certain quality of water being used as a criterion for

measures against discharges. This would result in discharges in different areas

being treated differently, which they maintain would disturb the balance of com-

petition. This premise is incorrect since environmental costs have to be con-

sidered as external costs. Ultimately the following solutions were agreed upon.

Given the principle that people should not be allowed to introduce the substances

on the black list into the environment, the aim of the measures is to bring a

gradual end to the discharge of these substances into the Rhine river basin. The

quality objective is in fact nil or the natural content. Internationally estab-

lished maximum discharge levels to be set by the Rhine Commission will be laid

down in a national licence system. The international levels will be based on

toxicity, persistence and bioaccumulation of the substances involved, and on the

best available technical means of reducing discharges. The discharge levels will

be regularly tightened. And the national licences will only be given for a re-

stricted period.

The Rhine Commission is responsible for ascertaining what effect the measures have

on the quality of Rhine water in addition to publishing the results of the mea-

surements. If necessary the Commission can propose other measures to reduce pol-

lution if it finds that the quality of the Rhine remains insufficient. Precisely

what these other measures involve is not described but they might f.i. be to pro-

hibit industries with production processes which cause severe water pollution from

setting up in the Rhine river basin or to take measures against other sources of

pollution.

As far as the grey list is concerned within two years of the convention coming in-

to force, the member states are to draw up programmes to reduce those discharges

which have a detrimental effect on the quality of Rhine water. Quality objectives


will have to be included in the programmes and time limits for implementing them.

Discharges of grey list substances will be permissible only with a licence in

which discharge levels determined on the basis of quality objectives are given.

The Rhine Commission will coordinate the national efforts with the aim that common

objectives will be reached.

I would plead in this context in favour of common quality objectives for inter-

national policy.

Arguments for quantified objectives for the quality of the surface water to be

reached, are that it gives certainty to downstream countries as well as upstream

countries. Together with industry the latter countries can aim their environmental

measures at achieving a clear situation. An even more important reason is that

quantified objectives provide environmental policy with the political base it

needs in daily negotiations with other interests.

A start has been made on developing quality objectives in the Common Market. It

means the same minimum objectives will be aimed at in the whole of the Common

Market, an important instrument not only for achieving the necessary quality of

the environment but also to prevent the environment being abused because of com-

petition. I shall return to this later.

I am happy that agreement was possible on a Rhine Chemical Convention. I am, how-

ever, aware that the definitive reply to the question of whether the treaty will

help to clean up the Rhine, will depend on many decisions which have yet to be

taken: which substances will be on the lists, which of them will have priority,

what specific emission standards will be agreed upon, and what quality objectives

will be set.

Much knowledge and perseverance will be needed in the coming phase, which will

first of all involve technical matters, if we are going to be successful. Again

and again the questions of to what extent health reasons and ecological require-

ments must limit human activities will be brought up.

The second international forum which I wish to discuss in some detail is the

Common Market.

The Common Market was set up in 1958 with primarily economic objectives. It is not

surprising that there are no provisions for environmental policy in it. The de-

cision of government leaders in 1972 to make a start on a European environmental

policy therefore created some legal problems. The unclear legal basis has meant

that the framework in which community decisions are taken is sometimes rather un-

intelligible. But that should not bother us too much.

More important is whether in the long run amendments to the Treaty will be needed

330 D. J. de Geer

if we are to be able to conduct an independent environmental policy at the

European level. In the current situation the absence of a specific legal basis

for environmental policy unduly reinforces the traditional economic objectives of

the Common Market.

European environmental policy started with an information agreement on national

provisions. Under this agreement the Community is given the opportunity to draw up

Community regulations within a maximum period of 10 months. Member states have

notified 148 draft regulations and 12 international agreements under this inform-

ation agreement of March 1973.

Meanwhile at the first "Environmental Council", a meeting of the Council of

Ministers of the Common Market in which ministers with responsibility for the en-

vironment are present, the first action programme on the environment was accepted.

The objectives and principles of environmental policy in the European communities

were established and a rather general programme was agreed upon. As far as

standards are concerned a few major points were already laid down. The member

states, depending on their situations, can set more stringent standards; European

environmental policy will establish minimum standards only. Another important

principle is that allowance will be made for health aspects as well as ecological

requirements.

In the four years which have since elapsed a number of important regulations on

water pollution have been drawn up. I would be doing the European Community an in-

justice if I were to give the impression that the directives which I am going to

list are the only ones; as part of the first action programme the European Com-

mission has submitted 36 proposals to the Council, 17 of which have led to de-

cisions by the Council to date.

In June 1975 an agreement was reached on the directive concerning the quality re-

quired for surface water intended for the abstraction of drinking water. In this

directive three categories of surface water apt for the production of drinking

water are fixed. From the worst quality an intensive treatment is necessary to

make drinking water. With normal treatment drinking water can be made from the

middle category and simple treatment is sufficient for the best one. In your

tables you will find the standards for the three categories and a more detailed

description of the three forms of treatment. The standards for each category are

divided between mandatory (I) and guide (G) standards.

As from July of this year in the community no drinking water may be made from

surface water of which the quality is worse than the a 3 standards. Temporary


exceptions are possible. Furthermore it is agreed upon that systematic plans of

action including time-tables shall be established for the improvement of surface

water. Over the next ten years considerable improvements have to be attained ac-

cording to the directive.

The European Commission intends to see to it that the improvement of the quality

of the surface water will increase the available reserves of surface water suit-

able for the production of drinking water.

In December 1975 a second directive in the same area followed. This one concerns

the quality of bathing water. The standards accepted under the directive - you

will also find them in your papers - have to be reached by member states within

10 years. The standards can be adapted to technical progress under a procedure

that gives opportunity to the European Commission to guarantee efficient

decision-making. The Commission will publicise a summary report on bathing water

in the member states and the most significant characteristics thereof.

This publication may help you to decide where to go during your summer holidays.

The European Commission made a proposal in August 1976 for quality objectives for

surface water in which fish can live. The proposed quality objectives are also

interesting from the ecological point of view. They are included in your papers.

In November 1976 the European Commission made a proposal for the quality of sur-

face water used for breeding shellfish.

These directives are still under consideration.

In May 1976 a directive on pollution caused by certain dangerous substances dis-

charged into the aquatic environment was accepted. In this directive common in-

struments for black and grey-list substances are dealt with. It took two council

meetings to agree on them. Like for the Rhine the choice between a policy based

on equal minimum emission standards all over Europe or equal minimum quality ob-

jectives had to be made. Again the solution roughly is that for the black-list

substances the emission standard system prevails and for the grey-list substances

the quality objective policy. At the special request of the U.K., for the black-

list substances a provision has been agreed upon that if certain international

quality objectives are reached, the country concerned is allowed not to apply the

common emission standards. The governments of the eight other common market coun-

tries declared that they will apply the directive without the escape clause.

The first emission-standards are in preparation for five substances on the black

list (mercury, cadmium, aldrin, dialdrin and andrin).

332 D. J. de Geer

In January and July 1975 the European Commission proposed two more directives to

do with the cleaning up of the paper pulp industry and the titan-dioxide industry.

These directives and the dangerous substances directive seem somewhat at odds with

each other, since these are industries which discharge primarily grey-list sub-

stances. Under the dangerous substances directive these discharges therefore come

under a policy in which the requirements are established depending on location.

Up till now no agreement on the directives could be reached. I am concerned what

the repercussions of a prolonged discussion may be for the further development of

European environmental policy. In the present precarious economic situation it is

risky to keep on fighting battles of principle. We might ask ourselves whether by

drawing up quality objective directives and working on the dangerous substances

directive we cannot go a long way towards creating a European water quality poli-

cy.

In the second action programme on the environment for the period up to 1981 much

attention has been paid to the prevention and reduction of pollution of fresh

water and seawater.

I have already indicated a number of points for possible further development at

European level. But we cannot avoid seeing these in the context of cultural con-

cerns. The cultural problem is political and probably unquantifiable. It is a

question of whether people will put up with living by polluted rivers and having

to swim in swimming baths. And it is also a matter of the risks which we are pre-

pared to take to see ecological balances which have been insufficiently analysed,

being fundamentally upset.

It is necessary to build a bridge between the cultural attitude and international

long term policy. Otherwise a sustained environmental policy on European scale

will be too easily disrupted by political considerations of the moment in indi-

vidual countries. Basic for such a long term policy seems to be an attitude of

thinking and acting for the future.

Margeret Mead in her book "Culture and Commitment" describes a culture in which

the future generation determines social behaviour. This as opposed to the past in

which parents determined it, and as a development of the present situation in

which the behaviour of contemporaries sets the tone. She thinks that the major

changes brought about by science and technology and the realisation that the

world can be completely destroyed will lead to this third culture.

Bearing this in mind it seems important to continue the development of environ-

mental standards and quality objectives which make allowance for the future. Our

technological era will be able to develop the necessary changes to production


processes and purification methods, to achieve the desired quality of the environ-

ment. The timing of their implementation will however be influenced by economic

considerations.

Such a set of standards and objectives can contribute to an international en-

vironmental policy which is politically feasible.

ANNEXE: THE RHINE BASIN

General Information

Rhine water is used for

(a) the abstraction of drinking water for human consumption,

(b) consumption by domestic and wild animals,

(c) the maintenance and fostering of the living conditions of wild animals and plants and the preservation of the self-purifying capacity of water,

(d) fishing,

(e) recreational purposes, bearing in mind hygienic and aesthetic requirements,

(f) the direct or indirect supply of fresh water to agricultural lands,

(g) the drawing off of water for industrial purposes, and for the need to pres-erve an acceptable quality of sea water.

(Enumeration taken from the 1976 Convention on the Protection of the Rhine against Chemical Pollution.)

Some figures on the utilisation of the Rhine

shipping: 120,000,000 tons (net)/year;

industry: 18,600,000,000 cubic meters/year (expected for 1980: this represents 27% of the average total dis-charge of the river);

drinking water: 793,000,000 cubic meters/year (expected for 1980);

discharge of waste: 20,000,000 tons/year.

Hydrology of the Rhine

In length (approximately 1000 kilometers) the Rhine is surpassed by many other rivers.

As a medium of transport it is matched only by the Mississippi and the St. Lawrence Seaway.

No other river has a larger concentration of population and industry than the Rhine.

The Rhine Basin — together with its tributaries1 basins, about 160,000 square km — contains approximately 30,000,000 inhabitants (average population density about 200 per square km).

Thanks to the fact that its sources are located in alpine regions and the fact that, compared to other rivers, its tributaries have a wide range of hydrological basins, the Rhine has a very balanced discharge.

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336 D. J. de Geer

CHEMICAL CONVENTION

Black List

Families and groups of substances

This list contains certain individual substances which belong to the following fam-ilies and groups of substances, selected mainly on the basis of their toxicity, per-sistence and bioaccumulation, with the exception of those which are biologically harmless or which are rapidly converted into substances which are biologically harm-less:

1. Organohalogen compounds and substances which may form such compounds in the aquatic environment.

2. Organophosphorus compounds.

3. Organotin compounds.

4. Substances in respect of which it has been proved that they possess carcino-genic properties in or via the aquatic environment.1

5. Mercury and its compounds.

6. Cadmium and its compounds.

7. Persistent mineral oils and hydrocarbons of petroleum origin.

Grey List

Families and groups of substances.

This list contains:

— substances belonging to the families and groups of substances in the Black List for which the limit values referred to in Article 5 of the Convention have not been determined,

— certain individual substances and categories of substances belonging to the families and groups of substances listed below, and which have a deleterious effect on the aquatic environment, which can, however, be confined to a given area and which depend on the characteristics and location of the water into which they are discharged.

Families and groups of substances referred to in the second indent,

1. The following metalloids and metals and their compounds:

1. 2. 3. 4. 5.

zinc copper nickel chromium lead

6. 7. 8. 9. 10.

selenium arsenic antimony molybdenum titanium

11. 12. 13. 14. 15.

tin barium beryllium boron uranium

16. 17. 18. 19. 20.

vanadium cobalt thalium tellurium silver

Where certain substances in the Grey List are carcinogenic, they are included in category 4 of this list.


2. Biocides and their derivatives not appearing in the Black List.

3. Substances which have a deleterious effect on the taste and/or smell of the products for human consumption derived from the aquatic environment and com-pounds liable to give rise to such substances in water.

4. Toxic or persistent organic compounds of silicon, and substances which may give rise to such compounds in water, excluding those which are biologically harmless or are rapidly converted in water into harmless substances.

5. Inorganic compounds of phosphorus and elemental phosphorus.

6. Non-persistent mineral oils and hydrocarbons of petroleum origin.

7. Cyanides, fluorides.

8. Substances which have an adverse effect on the oxygen balance, particularly ammonia, nitrites.

Continuous Biomonitoring Systems for Detection of Toxic LeveL· of Water Pollutants

J. H. KOEMAN*, C. L. M. POELS** and W. SLOOFF***

*Dept. of Toxicology, Agricultural University, DeDreijen 12, Wageningen, The Netherlands

** Testing and Research Institute of the Netherlands Water Undertakings, Rijswijk, The Netherlands

***National Institute for Water Supply, Voorburg, The Netherlands

ABSTRACT

The most important continuous biomonitoring systems which are presently being used for the assessment of water quality are discussed briefly. Special attention is given to their indicator value as well as to observation methods which may expand their perceptive faculty. It is concluded that a proper water quality assessment can not rely only upon the use of biological monitoring systems. They should be seconded by adequate multidetectional chemical monitoring methods.

INTRODUCTION

In recent years many attempts have been made to introduce bioassay models which will enable a more or less continuous surveillance of water quality with respect to the possible presence of potentially toxic pollutants. Most of the model systems consist of flow-through tanks in which aquatic organisms are exposed to surface water or industrial and domestic effluents; a certain biological response of the organisms is used as an indication for a change in water quality. Formerly water quality assessments were mainly based on the results of chemical analyses which were then compared with the available information about the toxicity of the compounds measured. The latter procedure has certain disadvantages, viz.: (1) In most cases chemical analyses of polluted waters will not yield a complete

picture of all compounds present, e.g. unknown toxic metabolites are overlooked completely.

(2) The chemical analyses are generally made discontinuously while there is often a delay between the time of sampling and the time the measurements are made.,

(3) The chemical detection of a compound in water not always implies that information is obtained about the biologically active form of the compound. For instance most metal analyses which at present are carried out routinely do not provide information on the species of the metal salts or ions involved. This implies that no accurate information is available about the 'biological strength' of the metal concentrations measured.

339

340 J. H. Koeman, C. L. M. Poels and W. Slooff

(4) The toxicological information may be incomplete, e.g. the toxicity for fish has not yet been studied in any detail.

(5) No investigations have been made showing how and to what extent the local environmental conditions like water temperature, hardness, pH, turbidity, oxygen content and salinity do affect the toxicity of the chemicals found.

In biomonitoring models the changes measured in the biological response of the test species used are likely to reflect a meaningful change in the chemical and/or physical conditions of the water concerned. Toxicological hazards measured by bioassay procedures may therefore be more realistic than those predicted from the results of chemical analyses and the available information on the toxicity of the compounds detected. However, the question can be raised whether the present continuous biomonitoring techniques for aquatic systems should be considered either as alternatives to the laborious conventional methods of water quality assessment or merely as valuable complementary tests. It is the aim of the present paper to discuss the value these tests may have for the qualitative and quantitative assessment of toxic pollutants in effluents and natural waters.

CONTINUOUS BIOMONITORING SYSTEMS

Various groups of aquatic organisms may serve as objects for continuous biomonitoring systems, including bacteria, algae, molluscs and fish. Up to date most systems use fish as bioassay organisms (e.g. Ref. 1, 2). Practical experience has been obtained with the following models:

Automatic Toxicity Monitoring Using Loss of Rheotaxis in Trout as Monitoring Principle

Trout in flowing water show a strong tendency to maintain their position when swimming against the stream (positive rheotaxis). When a toxic compound is introduced into the watersupply of the system two types of reaction may occur (1) the fish show an evading reaction and move downstream and (2) the compound affects the condition of the fish in such a manner that they are not longer able to maintain their position in the stream. The downstream movement of the fish can be used for the indication of toxic substances. Some systems are automated by using photoelectric cells (Ref. 3, 4, 5).

Registration of Breathing Patterns in Fish

In the most commonly used system a continuous registration is made of the opercular rhythm of fish (Ref. 6, 7, 8). The respiratory pattern may increase in response to changing environmental conditions. In case the breathing rate of the fish exceeds a predetermined limit which is supposed to be indicative of a toxic condition alarm circuits may be triggered. A related monitoring principle in the measurement of the socalled cough response (Ref. 9).

Registration of Locomotor Response of Fish

In this system a registration is made of the fish movement patterns by means of photo-electric monitoring devices (Ref. 10, 11). The response patterns are quantified and changes detected by following the movement activities during exposure to effluents or surface water. The studies reported sofar mainly refer to experiments where known toxic compounds have been tested under controlled conditions.

Continuous Biomonitoring Systems 341

The Rotatory-flow Technique

Fish are placed in a narrow socalled rotational tube through which the test water is flown and which is revolving with increasing velocity around the direction of flow. One measures the rate of rotation at which the fish is just unable to maintain the upright position. This critical response is used as a measure of fitness (Ref. 12, 13, 14).

THE INDICATOR VALUE OF THE PRESENT CONTINUOUS BIOMONITORING SYSTEMS

The monitoring systems described briefly above in general react to toxic compounds at concentrations well below the lethal or median lethal concentrations. Morgan (15) demonstrated for a number of compounds, including some metal ions (Cu2+, Cd2 + , Hg2 + , Pb2+) and organic (e.g. phenol, pentachlorophenol, parathion) and inorganic chemicals (cyanide, ammonia) that the opercular rhythm response in fish may registrate concentrations varying from 5 to 30 per cent of the 48-h LC50. Similar results were obtained by Slooff (16), who found values ranging from 1 to 30 per cent of the 48-h LC50 for acrylonitrile, Cd

++, Cu++, chloroform, cyanide, o- and m-dichlorobenzene, γ-hexachlorocyclohexane, pentachlorophenol, toluene, trichloroethylene and xylene. The experiments with the rheotaxis model show that the loss of positive rheotaxis may occur much earlier than death (Ref. 4, 17). Experiments with methylmercurie compounds and some metal ions further indicate that the cough response and the critical revolutions per minute parameter in the rotatory-flow model represent relatively sensitive indicators for a toxic action of the compounds concerned. The results obtained sofar support the conclusion that the systems mentioned can be used as early sensors or early warning systems for a fairly large group of chemical compounds. However, it should be recognised that the absence of a response in any of the systems concerned not necessarily implies that the water quality is acceptable from a general environmental or human health point of view. The systems are likely to detect mainly those chemical compounds which act more or less directly on the respiratory system and locomotor functions in relative high concentrations. Table 1 lists a number of chemicals known to induce a quick response in continuous biomonitoring systems or which are likely to do so on the basis of their mode of action.

On the other hand there are many groups of chemicals which are not likely to cause a more or less immediate or short term effect on the test organisms. Hazardous quantities of these may pass unnoticed. Examples can be found for instance among those chemicals which require bioactivation into reactive metabolites before they exert a toxic action. It is known that the activity of the important bioactivating microsomal mono-oxygenases is extremely low in trout and other fish species (e.g. 18, 19). There are strong indications that a number of halogenated aromatic compounds like PCB's, HCB and PBB!s become more toxic after certain bioactivation steps, which may at least partly explain their relatively low acute toxicity to fish. A second group of potentially hazardous chemicals concerns those of which the toxic action manifests itself only or mainly after chronic exposure. Examples can be found among the mutagenic and/or carcinogenic compounds with a relatively low acute toxicity to fish, like the polyaromatic hydrocarbons and aromatic amines. It is obvious that some of the compounds in Table 1 may also have a chronic toxic action at levels well below those detectable by means of the current biomonitoring models. Cairns et al. (20) demonstrated that the minimum sublethal zinc concentration detectable by means of a fish breathing response system amounted to 2-3 mg/1, whereas a reproduction study with the same species (blue gill sunfish) showed that 0.24 mg/1 was not safe for chronic exposure.


TABLE 1 Chemicals Known or Likely to Induce a Quick Response in Present Continuous Biomonitoring Systems

Chemical Mode of action

organic compounds

phenols

fish toxic pesticides like endrin, endofulfan, DDT,pyrethroids

organic solvents like benzene, toluene, alcohols, aliphatic halogenated hydrocarbons, ethers, glycols

methylmercury compounds

aromatic nitro and amino compounds

organic nitriles, e.g. acetonitrile

direct action on gill membranes, also uncoupling of oxydative phosphorylation (PCP)

effects on nervous system (e.g. sensory and motor neurons, lateral line organ)

narcotic action and irritating action on gills

effect on central nervous system

interference with respiration (e.g. methomoglobinemia)

irritation of mucous membranes and inhibition of cellular respiration via CN~

inorganic compounds

2 + 2 + free metal ions such as Hg , Cd , Cu2+, Zn2+, CrO^2"

ammonia

nitrite

cyanide, sulphides

direct action on gill membranes

interference with physiology of gills (gasexchange, osmoregulation)

idem, and induction of methemoglobinemia

inhibition of cellular respiration

As a third group of relatively non-reactive but ultimately hazardous chemicals one may refer to compounds which are not available for uptake by fish at the time of exposure. This applies in particular to compounds which become readily adsorbed to suspended organic material and clay minerals such as metal salts and organochlorine pesticides. This implies that solutions and more or less free suspensions of these chemicals may be recorded while the same concentration in particle bound form will not show any effect. The former condition is likely to


occur mainly after obvious spills of the concentrated products like pesticide formulations and highly contaminated industrial effluents. The latter condition generally reflects a gradual release originating from a multitude of normal usages in industrial, agricultural and domestic applications.

Considering the limited screening ability of the short term test systems concerned the conclusion can be drawn that these can certainly not replace the conventional methods of water quality assessment based on chemical analyses and the predictive evaluation of toxicological risks. However, they may represent valuable complementary tests under certain circumstances. The most successful application seems to be monitoring of surface waters at strategic points, e.g. localities in the neighbourhood of industrial centers, and industrial and domestic effluents. For most natural waters the continuous operation of the present type of biomonitoring systems seems to be less rewarding because of the reasons referred to before. For instance continuous monitoring of the Rhine for the past period of 3^ years by means of a sensitive rheotaxis model did not once yield an alarm level. Apparently acutely toxic concentrations of toxicants did not occur in this period, which was further confirmed by the absence of fish kills not due to low oxygen content and the results of chemical analyses. Still many toxic chemicals were found to be present and occasionally in potentially hazardous amounts. However, it certainly is possible to expand the perceptive faculty of biomonitoring systems for natural water by increasing the number of health parameters which is checked systematically in order to detect any adverse change in test organisms. This will generally imply that additional tanks for long term-observations will have to be operated next to the flow-through systems used for the short term warning systems.

NEWER DEVELOPMENTS IN CONTINUOUS BIOMONITORING SYSTEMS

A perfect biomonitoring system for water quality assessment should be able to detect a wide range of acute as well as chronic toxicological hazards. Preferrably some indication should also be obtained about the type of chemicals involved. In recent years various attempts have been made to expand the number of physical parameters measured in fish and other organisms in long term biomonitoring systems. A few examples of meaningful parameters which can be measured in combination with continuous flow systems are summarized in Table 2. Most of these refer to parameters commonly checked in routine toxicity studies. Poels and Strik (21) reported body weight, organ weight and hematological measurements in trout continuously exposed to Rhinewater for 9 months. A control group was simulataneously exposed to groundwater. A significant increase in liver weight was notable after 3 months in fish kept in Rhinewater, which may indicate that certain enzymes like the mixed function oxydase systems were induced by enzyme inducing compounds known to occur in this river, e.g. PCBTs and HCB. In the meantime it has been confirmed by direct measurement, that Rhinewater causes an increase in mixed function oxydase activity in the livers of rainbow trouts (Salmo ga-irdneri) while the activity of certain other enzymes was also affected, e.g. liver GPT, LDH and alkaline phosphatase (22). The inhibition of acetylcholinesterases and related enzymes like aliesterases may serve as indicator for hazardous or near hazardous levels of organophosphorus compounds and carbamates (see for instance references in 23). Haematological parameters like blood glucose, haemoglobin and haematocrit values may also represent valuable parameters (e.g. 24, 25). It is most likely that more enzyme systems and clinicochemical parameters can be used as appropriate quality guides for monitoring changes in the chemical composition of the aquatic environment. Activated microbial mutagenicity tests like those devised by Ames et al. (26)


have been applied to surface waters in order to detect the possible presence of mutagenic and/or carcinogenic compounds (e.g. 27). Furthermore attempts are made to use certain species of fish as cytogenetics models for the detection of compounds causing genetic effects. Kligerman et al. (28) drew the attention to the central mudminnow (Umbra Ιϊητί) which seems to be a useful test organism because it has a relatively low number of fairly large chromosomes. Recently it was shown by Prein et al. (29) that eastern mudminnows (Umbra pymaea) which possess a similar karyotype showed a significant increase in chromosome breaks a after continuous exposure to Rhinewater for periods up to 11 days. Studies on the mutagenicity of Rhinewater extracts by means of a bacterial mutagenicity test system revealed that the aromatic hydrocarbon fraction produced a positive mutagenic response. This may indicate that one or more of the compounds present in the latter fraction were also responsible for the cytogenetic changes found in fish.

TABLE 2 Examples of Physical Parameters Measurable in Continuous Biomonitoring Systems

Parameter Indicator for:

drug enzyme systems in fish (e.g. mixed function oxydases)

drug enzyme inducers (e.g. halogenated hydrocarbons)

brain acetylcholinesterase in fish

haematological assessments, e.g. blood glucose, haemoglobin and haematocrit values

organophosphorus compounds, carbamates

chemical or other stress factors

mutation frequency change in activated microbial mutagenicity tests after exposure to water or organ extracts

mutagenic and/or carcinogenic compounds

chromosomal aberrations in gill cells of fish (e.g. mudminnows)

mutagenic and/or carcinogenic compounds

egg production, hatchability and survival of fry in reproducing fish species

chemicals interfering with reproduction

gross pathological and histopathological examinations after short and long term exposure

agents causing gill lesions and hepatotoxic, nephrotoxic tumorigenic and other more or less specific types of effects

Observations on reproduction can also be carried out in biomonitoring systems by using suitable indicator species such as zebra fish (Braohydanio rerio) (e.g. 30) or trout.

Finally samples of indicator organisms can be submitted to gross pathological and/or histopathological examinations. Lesion in organs such as gills, liver and


kidney may be indicative for the presence of certain toxic agents. For instance certain changes in the structure of the gills like a proliferation of epithelial cells in gill lamellae may indicate that the fish are exposed to toxic sublethal levels of metals (e.g. chromium salts, Ref. 31) or other compounds (ammonia, phenols). Early signs of spinal and vertebral aberrations, which may be indicative for compounds such as Zn++, Cd++, toxaphene and certain organophosphorus compounds can be detected only after dissection of the fish or with the aid of radiography (e.g. Ref. 32).

These additional observation methods, which are already being used in a number of existing long term biomonitoring programmes in various countries, may add valuable complementary information to that obtained through the observation methods described before. Certainly much less chemicals will pass the system without yielding a response than in the rheotaxis, opercular rhythm and other models. However, this does not yet warrant the conclusion that we are now in a position to devise monitoring systems capable of detecting all or at least most toxicological hazards which may turn up in surface waters or industrial and domestic effluents. Still there are many gaps in our knowledge about the modes of action by which known chemicals may induce toxic effects in living organisms. Therefore the symptoms seen in the indicator organisms used in biological monitoring models will in many cases not indicate clearly what the nature is of the factors by which they are caused. Moreover a number of potentially hazardous compounds may still escape our attention. Hence, it may be concluded that as long as aquatic pollution with many known and unknown chemical compounds continues at the present level the assessment of water quality can not rely only upon these biological monitoring models. They should be seconded by adequate multidetectional chemical monitoring methods.

REFERENCES

(1) G.E. Glass, (Editor) (1973) Bioassay Techniques and Environmental Chemistry, Ann Arbor Science Publishers, Michigan.

(2) J. Cairns, K.L. Dickson and G.F. Westlake (Editors) (1977) Biological Monitoring of Water and Effluent Quality, American Society for Testing and Materials, Philadelphia.

(3) C.L.M. Poels, Continuous automatic monitoring of surface water with fish, Water Treatment and Examination 24, 46 (1975).

(4) C.L.M. Poels, An automated system for rapid detection of acute high concentrations of toxic substances in surface water using trout, in Cairns et al. (see Ref. 2 ) , pp. 85-95.

(5) W.K. Besch, A. Kemball, K. Meijer-Waarden, and B. Scharf, A biological monitoring system employing rheotaxis of fish, in Cairns et al (see Ref. 2), pp. 56-74.

(6) W. Spoor, T.W. Neiheisel, and R.A. Drummond, An electrode chamber for recording respiratory and other movements of free swimming animals, Trans. Am. Fish Soc. 100, 22 (1971).

(7) W.S.G. Morgan, An electronic system to monitor the effects of changes in water quality on fish opercular rhythms, in Cairns et al. (see Ref. 2), pp. 38-55.

(8) W. Slooff and B.C.J. Zoeteman, Toxicological aspects of some frequently detected organic compounds in drinking water, RID-report 76-15. Report prepared for the Commission of the European Communities. Brussels (1976).

(9) R.A. Drummond, G.F. Olson, and A.R. Batterman, Cough response and uptake of mercury by brook trout, Satvelinus fontinalis 9 exposed to mercuric compounds at different hydrogen-ion concentrations, Trans. Amer. Fish. Soc. 103, 244 (1974).


(10) K. Müller and K. Schreiber, Eine Methode zur Messung der lokomotorischen Aktivität von Süsswasserfischen, Oikos 18, 135 (1967).

(11) H. Kleerekoper, Some monitoring and analytical techniques for the study of locomotor responses of fish to environmental variables, in Cairns et al. (see Ref. 2), pp. 110-120.

(12) P.E. Lindahl and E. Schwambom, A method for the detection and quantitative estimation of sublethal poisoning in living fish, Qikos 22, 210 (1971).

(13) P.E. Lindahl, S. Olofsson, and E. Schwambom, Improved rotatory-flow technique applied to cod (Gadus morrhua L.)9 Water Research 10, 833 (1976).

(14) P.E. Lindahl, S. Olofsson, and E. Schwambom, Rotatory-flow technique for testing fitness of fish, in Cairns et al. (see Ref. 2), pp. 75-84.

(15) W.S.G. Morgan, An electronic system to monitor the effects of changes in water quality on fish opercular rhythms, in Cairns et al (see Ref. 2), pp. 38-55.

(16) W. Slooff, National Institute for Water Supply, Voorburg, The Netherlands, unpublished data.

(17) C.L.M. Poels, Flow-through systems for the continuous monitoring of raw water using trout. Proceedings 11th Congress IWSA, Amsterdam 13-17 Sept. 1976, pp. 010-015.

(18) J.H. Dewaide (1971) Metabolism of Xenobiotics. Thesis. University of Nijmegen, The Netherlands.

(19) C.H. Walker, The significance of pesticide residues in the environment, Outlook on Agriculture 9, 16 (1976).

(20) J.J. Cairns, R.E. Sparks, and W.T. Waller. The relation between continuous biological monitoring and water quality standards for chronic exposure. Bioassay techn. and environm. chemistry, 383-401 (1973).

(21) C.L.M. Poels and J.J.T.W.A. Strik (1975), Chronic toxic effects of the water of the river Rhine upon rainbow trouts, in Sublethal Effects of Toxic Chemicals on Aquatic Animals (J.H. Koeman and J.J.T.W.A. Strik editors). Elsevier, Amsterdam, Oxford, New York, pp. 81-91.

(22) C.L.M. Poels, see communication elsewhere in the present proceedings. (23) W.A. Thomas, G. Goldstein and W.H. Wilcox (1973) Biological Indicators of

Environmental Quality. Ann Arbor Science Publishers, Michigan. (24) P.C. Blaxhall (1972) The haematological assessment of the health of

freshwater fish. J. Fish. Biol. 4, 593. (25) E.K. Silbergeld (1974) Blood glucose: a sensitive indicator of

environmental stress in fish. Bull. Environ. Cont. Toxicol. 11, 20. (26) B.N. Ames, J. McCann, and E. Yamasaki, Methods for detecting carcinogens

and mutagens with the SaImonella/Microsome mutagenicity test. Mutation Research 31, 347 (1975).

(27) W. Pelon, B.F. Whitman and T.W. Beasley, Reversion of histidine-dependent mutant strains of Salmonella typhimurium by Mississippi river water samples. Environm. Sei. Techn. U-, 619 (1977).

(28) A.D. Kligerman, S.E. Bloom and W.M. Howell, Umbra limi: a model for the study of chromosome aberrations in fish. Mutation Research 31, 225 (1975).

(29) A.E. Prein, G.M. Thie, G.M. Alink, J.H. Koeman and C.L.M. Poels, Cytogenetic changes in fish exposed to Rhinewater, Manuscript.

(30) J.E. Kihlström, C. Lundbey and L. Heelth, Number of eggs and young, produced by zebrafishes (Brachidanio rerio, Ham.-Buch.) spawning in water containing small amounts of phenylmercurie acetate. Environmental Research 4, 355 (1971).

(31) J.J.T.W.A. Strik, H.H. de Iongh, J.W.A. van Rijn van Alkemade and T.P. Wuite (1975), Toxicity of chromium (VI) in fish, with special reference to organoweights, liver and plasma enzyme activities, blood parameters


and histological alterations,in Sublethal Effects of Toxic Chemicals on Aquatic Animals (J.H. Koeman and J.J.T.W.A. Strik, editors). Elsevier. Amsterdam, Oxford, New York, pp. 31-41.

(32) Bengt-Erik Bengtsson (1975), Vertebral damage in fish induced by pollutants, in Sublethal Effects of Toxic Chemicals on Aquatic Animals (J.H. Koeman and J.J.T.W.A. Strik, editors). Elsevier. Amsterdam, Oxford, New York, pp. 23-30.

Feral Aquatic Organisms as Indicators of Waterborne Environmental Carcinogens

R. A. SONSTEGARD

Department of Microbiology, University ofGuelph, Guelph, Ontario, Canada

ABSTRACT

Our aquatic environments are fouled with chemicals and classes of chemicals which have carcinogenic potential. In this paper, the potential utility of monitoring aquatic organisms as a sentinel system for the early detection and identification of waterborne carcinogens is described.

INTRODUCTION

Within the next fifty years, according to some forecasts, the population of the U.S.A. will double, and the demand for water by cities, industry, and agriculture will grow faster than the population. Today, these water uses add up to some-thing like 350 billion gallons a day (BGD), but by 1980, by some estimates, they will amount to 600 BGD. By the year 2000, demand for water is expected to reach 1000 BGD, considerably exceeding the essentially unchanging supply of dependable fresh water, which is estimated at 650 BGD (1). More and more water will have to be reused, and the public health hazards of water reuse remain largely unknown.

It has been documented that our aquatic environments are fouled with chemicals and classes of chemicals which have carcinogenic potential (2-10). Although there is little evidence that carcinogens in water have produced widespread cancer problems in man, it is not difficult to envisage such a possibility. Presently, because of ground water depletion, many cities are processing for drinking purposes, water which may be in excess of 30% recycled water (11). Many inorganic and organic compounds (some known carcinogens) are not removed in current water treatment facilities (11, 12). In fact, treatment may be producing carcinogens (i.e. chlorination may produce chloroform and carbon tetrachloride, both known carcinogens) (13-15).

Of particular concern with regards to the above are the reports (16-18) on the possible implications of cancer-causing substances in Mississippi River water. These studies present presumptive epidemiological evidence which suggests a significant relationship between cancer mortality in white males drinking water which was obtained from the Mississippi River in the New Orleans area. The reports strongly suggest that drinking water from the Mississippi River is causally related to cancer mortality in more than one million persons in Louisiana who depend on that source for their drinking water supply. Although the results

349

350 R. A. Sonstegard

of the New Orleans study cannot as yet be considered as conclusive evidence that cancer is in fact being caused by consuming contaminated water, these very suggestive findings must be fully taken into consideration.

It is apparent that there is an urgent need for new model systems to evaluate our aquatic environments, particularly with regards to carcinogens. Although much attention has been given recently to the potential health problems of waterborne carcinogens (18), there have been few innovations and/or studies initiated for monitoring for these agents in aquatic environments. In this paper we report several approaches to this problem using aquatic organisms as indicators of aquatic pollutents.

ENVIRONMENTAL SURVEILLANCE OF NEOPLASMS IN AQUATIC ORGANISMS

One of the most productive areas in human cancer research has been epidemio-logical studies. The standard method in cancer epidemiology is to tabulate incidence data for each histologic type of neoplasm in man and to compare these data amongst various populations over the world. In this way, cancer incidence maps can be made, and it is then possible to establish correlations with industrial activities, radiation exposure, etc. However, in only a few cases which mainly include tumors in industrial workers (i.e. vinyl chloride) has it been possible to identify the oncogenic compounds.

As a result of our mushrooming technological society, thousands of new compounds are annually being released into our environment. Few of these compounds (i.e. 5,000 up to 1972) have been screened for carcinogenic activity <Ί9). This represents a small number of the tens of thousands, perhaps hundreds of thousands, of chemicals which have extensive environmental contact with man, and ought to be screened for carcinogenic activity.

Very likely, in lieu of the above, the incidence of neoplasia in man could increase because of new carcinogens, increased levels of carcinogens already in the environment, or the combined effects of the same which might be revealed decades after their introduction. Furthermore, environmental carcinogens have a tendency to express themselves in human beings some 30 years post exposure.

The aquatic environment is one of the ultimate recipients of man's environmental pollution. The organisms inhabiting it are subject to the same or even more concentrated chemical and biochemical pollutants as man. Therefore, aquatic organisms may turn out to be sensitive indicators of what the future holds for man in terms of waterborne carcinogens.

It is apparent that there is an urgent need for a new model systems to evaluate our aquatic environments, particularly with regards to carcinogens. In the following, we summarize several investigations which have explored the potential utility of monitoring aquatic organisms as indicators of waterborne carcinogens. As free-living organisms which co-inhabit man's biosphere, they may act as a bridge between laboratory testing and epidemiological surveillance of human populations.

In this investigator's laboratory we have made extensive tumour frequency studies in feral fishes inhabiting the Great Lakes of North America (Superior, Michigan, Huron, Erie, and Ontario). During the course of these investigations (over 50,000 fishes were captured and necropsied) and epizootics of neoplasms in the following species were discovered which appeared to have an environmental etiology (20-26).

Feral Aquatic Organisms 351

Gonadal tumours in cyprinid fishes (carp, Cyprinus carpio; goldfish, Cassarius auratus; and carp X goldfish hybrids) were found with tumour frequencies approaching 100% in specific age-sex groups. Epizootiological studies and reviews of museum collections (made prior to 1952) suggests that the tumour is being induced by environmental factor(s) being discharged into the lake system.

Epizootics of skin tumours (papillomas) were found in white suckers (Catastomus commersoni) throughout the lake system. However, elevated tumour frequencies (i.e. 50.8%) were found clustered around the industrial-urban complex on Lake Ontario. In addition to the clustering recorded, there was an anatomic shift in the anatomical location of the tumour to a site (lips) which has near-constant contact with bottom sediments (white sucker is a bottom dwelling species). The anatomical shift in the location of the tumour, together with the clustering recorded strongly suggest that factor(s) in the bottom sediments are enhancing the oncogenicity of a C-type virus associated with the neoplasm.

Epizootics of thyroid hyperplasia (goiter) were found in Great Lakes coho salmon (Oncorhynchus kisutch) with frequencies of occurrence as high as 79.5% recorded. Epizootiological studies conducted on the lake system (1972-1976) indicate that the frequency of occurrence of the goiters to be increasing, and that environ-mental goiterogens (possibly pollutants) may be involved in the etiology of the thyroid dysfunction.

Brown and co-workers (27) reported field epizootiological studies of the fre-quency of occurrence of neoplasia in fishes inhabiting a heavily polluted river in the U.S.A. using a Canadian watershed as a control. They reported elevated tumour frequencies in fishes captured from polluted waters (4.38%) compared to the control watershed (1.03%) suggesting a possible environmental role.

On the coastal regions of the Pacific Ocean, a number of investigations have been made of epizootics of papillomas in pleuronectid fishes (28-32). These fish live under a variety of conditions in both polluted and non-polluted waters, often in close proximity to human habitation and industries. The papilloma is readily detected in superficial examination of captured fishes, and as such provides an ideal marker for field epizootiological studies. To date, no definitive correlations have been made with regards to elevated tumour frequencies in polluted vs. non-polluted waters, however, with the advent of the trans-Alaska pipeline, the problem of incidence of neoplasia in these fishes takes on new relevance. It is apparent that the disease has the potential of becoming a valuable environmental barometer for marine pollution.

Epizootics of neoplastic or proliferative diseases have been reported in marine shellfish from a number of zoogeographical regions (33-36). Although the etiology of the disease in shellfish remains obscure, some authors have implicated various chemicals, particularly those present in fossil fuels or coal-tar derivatives (35). Dunn and Stich (37) have reported the association of the carcinogen benzo(a)pyrene from environmentally contaminated shellfish which suggests that these animals might be good indicators of this group of compounds. As an environmental sentinel animal for the detection of carcinogens, shellfish have the advantage in that they are non-migratory, and as such might be able to "pin-point" cancer causing pollution.

Rose (38) reported epizootics of tumourous growths in reptiles (tiger sala-manders, Ambystoma tigrinum) which were captured in lagoons receiving treated sewage effluents. The salamanders affected were in the larvae stage, having spent their entire lives in the sewage lagoons. To date, the tumourous growths


have been found only in the lagoons receiving sewage, suggesting that the neo-plasms in this species may be induced by factors in the water. Although the etiology of the neoplasms in the salamanders is unknown, these fascinating observations strongly suggest extrinsic environmental agents (possibly pollutants) play a role in the etiology of the disease.

Probably the best example of the potential utility of monitoring aquatic organisms to detect environmental carcinogens are investigations linking industrial discharges and the occurrence of tumour-like growths in red algae (Porphyra tenera) in Japan where it is cultivated for commercial purposes (39-41). The tumour-like growths were found in several locations on the shoreline of Japan, although a very high incidence of algal growths were found in the vicinity of an industrial complex. Mud samples from this area induced similar growths in Porphyra within 3 weeks under laboratory conditions. Chemical analyses of the mud and subsequent laboratory testing revealed that the inducing agents were polycyclic aromatic hydrocarbons including benzanthrone, dibenzan-throne, and 2-chloroanthraquinone. Thus, the source and identity of cancer-causing pollution was pin-pointed to the coal industry by monitoring the environ-ment. The potential utility of propogating the algae in mud samples from a variety of aquatic environments is obvious.

The above examples of proliferative and/or neoplastic conditions in fishes, shellfish, reptiles, and algae inhabiting both marine and freshwater environ-ments all support an environmental etiology. Although much work needs to be done concerning qualitative and quantitative aspects of the role of extrinsic environmental factors in the induction of these proliferative conditions, the value of capitalizing on these animals as sentinels for the early detection of waterborne carcinogens is apparent. Monitoring of both feral and caged aquatic organisms placed in a variety of aquatic environments may well serve to provide increasingly needed environmental indices.

BIOCONCENTRATION OF AQUATIC POLLUTENTS BY AQUATIC ORGANISMS

One of the major technical problems in the chemical analysis of water for car-cinogens is that most aquatic pollutents are highly diluted. The ability of aquatic carcinogens to concentrate persistent organic compounds and heavy metals is almost legendary, a trait which had led to the detection and identification of a number of aquatic pollutents. For example, polychlorinated biphenyls (PCB*s) levels in tissues of adult Great Lakes coho salmon are nearly five logs higher than those found in water, which has led to these fish being banned for commercial sale. In this way, chemical analysis of tissues for the presence of pollutents has provided a valuable adjunct in environmental studies and has played an important role in the protection of consumers from carcinogens in foods.

In addition to concentrating compounds from aquatic environments facilitating chemical analysis as discussed above, the bioconcentration process provides a potential avenue of carcinogen assessment. For example, fat depots accumulate lipophilic organics (i.e. PCB) and bile from fishes has been shown to be a site of bioconcentration of xenobiotics (42). It would seem likely, in lieu of the above, that in vitro carcinogen-mutagen testing procedures (i.e. Ames test, virus enhancement, and DNA repair) performed on extracts prepared from these sites might give a qualitative assessment of waterborne carcinogens. This also suggests that water quality with particular reference to carcinogens, might be monitored for with caged aquatic organisms thereby pin-pointing the source of cancer causing pollution.


ENZYME INDUCTION IN AQUATIC ORGANISMS

Metabolic activation is necessary for most chemical carcinogens, the exception being compounds sufficiently reactive to interact with cellular macromolecules by themselves. Activation is usually mediated by microsomal hydroxylases, the induction of tumours being dependent on the activation reactions.

Payne (43) reported that the liver of a marine fish (cunners, Tautogolubrus adsperus) has an inducible benzopyrene hydroxylase (benzypyrene is a carcinogen), the enzyme induced in the species by exposure to petroleum. Payne (43) made exploratory field epizootiological studies of feral cunners collected from petroleum contaminated and control sites in the Atlantic Ocean. In these studies, he found elevated levels of benzopyrene hydroxylase activity in the livers of cunners captured in the petrochemical polluted waters. This study suggests that the presence of the enzyme in a given species might provide a practical "built-in" monitor for aquatic petroleum pollution.

With regards to metabolic activation of precarcinogens in fishes, Stich and Acton (31) reported that S-9 fractions prepared from fish (trout and flounder) gave good metabolic activation with several known precarcinogens. This test system may be a particularly valuable probe in assessing the relative sensitivity of various aquatic organisms to known environmental carcinogens, and as such, may be proved important indiced in establishing epizootiological pointers for laboratory cause-effect studies in a given species.

AQUATIC ORGANISMS AS LABORATORY TEST ANIMALS

The potential value that aquatic organisms might play as laboratory test animals for the detection of carcinogens is perhaps best exemplified by the aflatoxin story. In the early 1960fs, trout hatcheries in the Pacific Northwest of the U.S.A. were plagued by epizootics of hepatomas. Field epizootiological studies established that mold contaminated feed was associated with these epizootics, and through biochemical and feeding trials it was discovered that the etiological agent was a metabolite (aflatoxin) produced by a common mold (Aspergillus flavus). On the molar basis, aflatoxins are the most powerful hepatocarcinogens known. Of the animal systems tested to date, rainbow trout are the most sensitive; hepatomas being induced with 0.05 ppb aflatoxin (44). Recently, there has been considerable speculation that the hepatocarcinomas in man in the Orient might be associated with aflatoxin contaminated diets (45). Although the aflatoxin feeding trials established that fish might be an unusually sensitive animal for carcinogen detection, little has been done to explore this potential. In a limited experiment,.Stanton (46) added the carcinogen diethyl-nitrosamine to aquarium water and induced hepatomas in an aquarium species (Brachydanio rerio).

It is apparent that the potential utility of aquatic organisms as an n vivo test system for carcinogen detection has been sorely overlooked. They have a potential advantage in water testing in that they may be held in aquaria in direct contact with the test chemical or perhaps even being held in cages directly in the environment.

DISCUSSION

It has been estimated that 60 to 90% of the cancers in man are environmentally induced (47). Furthermore, in man, environmental carcinogens have a tendency to express themselves about 30 years post exposure. Waiting for human evidence of environmental carcinogens means acceptance that prevention can only be imple-


mented posteriori, and that "human experimentation" will continue until over-whelming evidence of a blatent cancer hazard imposes adoption of radical measures. The New Orleans "experience" gives urgency to the question of water-borne carcinogens and suggests a dire need to detect and identify factors in the aquatic environments that could be contributors to the etiology of cancer in man.

At present, the screening of compounds for carcinogenic potential is pro-gressing at a rapidly accelerated rate due to advances in the development of "short term" tests (48). Carcinogen testing, however, is largely confined to testing of purified chemicals in a battery of in vitro and in vivo assays. It is increasingly apparent, however, that evaluation of the carcinogenic potential of chemicals found in the aquatic environments is very complex. Even if the specific carcinogenic potential of a defined industrial or agricultural chemical is established, the breakdown products in the environment might be even more difficult to assess. Natural biodegradation or specific treatment processes (i.e. chlorination) of industrial and domestic effluents may lead to the development of new compounds quite different from that of the present compound. Another factor that confounds evaluation of environmental agents is the com-bined and/or synergistic effects that may result from a mixture of carcinogenic and noncarcinogenic compounds.

Today, science and technology are in a transition period. The greater pro-portion of cancers in industrial nations are related directly or indirectly to factors which have been in the environment a long time. Society is clearly entering a new phase, where exposure to a large number of hitherto unrecognized chemical hazards is possible. Since testing and rapid screening procedures are imperfect, some method should be established to provide epidemiological sur-veillance (in addition to humans) on a global scale. This would permit monitoring of cancer patterns and exposures in different environments. It would facilitate prompt investigations of new hazards, thus avoiding the necessity of setting up emergency axl hoc studies after an unexpected hazard has been identified, due to the absence of human data.

In this paper, we have discussed several avenues of investigations which have explored the potential utility of monitoring feral aquatic organisms as indicators of waterborne carcinogens. Inasmuch as they co-inhabit man's bio-sphere, they may act as a bridge between laboratory testing and epidemiological surveillance of human populations. While monitoring for cancer problems in aquatic organisms may be highly attractive, there are logistical problems. There exists a need to standardize methods of assessing environmental exposure, epizootiological parameters (i.e. age, sex), and correlation studies to control sites. In addition, a great deal needs to be known about the natural biology of the organism-neoplasm being monitored.

A cancer registry exists for tumours in lower animals which should prove to be a valuable adjunct in the assessment of regional epizootiological studies (49). Inclusion of data in such a registry on a global basis may pin-point the associationship between environmental exposure and neoplasm development. This approach might lead to the early detection and elimination of carcinogens in aquatic environments and may eventually yield an extraordinary impact in diminishing the incidence of cancer in man.

R.A, Sonstegard, Research scholar of the National Cancer Institute of Canada. This work was supported by grants from the National Cancer Institute of Canada, Environment Canada, and the Ontario Ministry of Natural Resources.


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Sensorily Perceptible Organic Pollutants in Drinking Water

B. C. J. ZOETEMAN, G. J. PIET andC. F. H. MORRA

National Institute for Water Supply, 2260 AD Leidschendam, The Netherfonds

ABSTRACT

An inquiry carried out in 1976 among 3073 adults in The Netherlands indicates that 6.9% of the population rate the taste of their drinking water as objectionable or worse, which value is 15% for the part of the population supplied by water derived from surface water. Sensory water quality assessment is a major factor in the overall water quality assessment by consumers as measured by the acceptability and reliability rating. Furthermore water taste affects the quantity of actually consumed drinking water. Assessment of taste and odour of 20 types of drinking water by a selected panel, consisting of 50 subjects, in combination with detailed chemical analysis of the water samples showed the major role in water taste impairment of organic micro-pollutants as compared to inorganic compounds. Among the 280 organic compounds detected in the studied drinking water tyeps, 8 organo-chlorine compounds, which were insufficiently removed during treatment from the contaminated Rhine water, could be identified as probable causes of impaired taste.

INTRODUCTION

The consumer of drinking water generally requires that the water is odourless, colourless and tasteless and has a temperature of approximately 10°C in order to obtain the refreshing effect during the quenching of thirst. An inquiry held in June 1976 in The Netherlands among 3073 individuals of 18 years and older, which was carried out by Mr. J.J.C. Karres of the Central Bureau of Statistics (Voorburg) in cooperation with the National Institute for Water Supply (Leidschendam) (1) showed that onlyO.. 2% of the persons rated water temperature as marked lukewarm or worse, 1.3% found water colour and turbidity objectionable or worse, 3.2% rated water odour as objectionable or very bad and 6.9% assessed water taste as objectionable or worse. So water taste is the most critical senso-rily perceptible aspect of drinking water quality.

It has been shown that "tasteless" water should contain a number of constituents in concentrations corresponding with those in saliva (2). Particularly inorganic compounds such as NaHC03 and CaS04 contribute to the pleasant neutral taste of drinking water when present in optimal concentrations (3).

The mentioned inquiry indicated that impaired water taste is mainly related to supplies which derive drinking water from contaminated surface water sources in

359

360 B. C. J. Zoetman, G. J. Piet and C. F. H. Morra

The Netherlands, as shown in table 1.

TABLE 1 Type of Water Source and Taste Assessment of Drinking Water in The Netherlands (n=3073)

Water source

groundwater surface water

% of persons scoring in a taste category

good

38.7 14.9

not perceptible

52.8 52.0

weak, not obj ectionable

5.9 18.5

object^ ionable

1.8 10.1

very bad

0.8 4.5

% of total

66.1 33.9

* including bankfiltered water mixed with groundwater

In this paper first the impact of sensory water quality assessment on consumer attitude will be discussed followed by an evaluation of the results of a study of 20 types of tapwater, in The Netherlands, relating to the chemical causes of impaired water taste and odour as assessed by a selected panel. Finally particular attention will be given to those organic micropollutants, dis-charged into the water of the Rhine, penetrating into the final drinking water and causing an impaired taste.

CONSUMER ATTITUDE AND SENSORY WATER QUALITY ASSESSMENT

Based on the results obtained from the mentioned inquiry it can be shown (1) that water taste is the major factor determining the judgement of consumers on the overall acceptability and reliability of water quality from a public health point of view. The correlation coefficients resulting from linear regression calculations for data relating to 50 communities in The Netherlands are summarized in table 2.

TABLE 2 Correlation between Acceptability and Reliability Ratings and Sensorily Perceptible Water Quality Aspects (n=1933)

Water Quality Aspect

Acceptability rating Reliability rating

Correlation coefficient(r) with

Taste rating

0.92 0.72

Colour rating

0.56 0.63

Ca-hardness

0.56 0.43

* r> 0.36 : a = 0.01 and r> 0.45 : a = 0.001

Apart from the reduced esthetic properties of drinking water, impaired taste and odour also affect the actually consumed quantity of water. This is illustrated by the data of table 3 which apply to individuals of 18 years and older. The total quantity of daily consumed tapwater in the form of tea, soup, coffee, water or lemonade, at house as well as elsewhere, amounts to .1.36 liter/head for those rating water taste as good and to 1.17 for those rating water taste as objectionable or worse. Table 3 further shows that the reduction of 0.19 1/head/day can be almost completely attributed to the reduced consumption of drinking water as such which value drops from 0.32 1/head/day to 0.15 1/head/day, which is a reduction of more than 50%.

Sensorily Perceptible Organic Pollutants 361

TABLE 3 Effect of Water Taste Rating on Total Water Consumption and Quantity of Drinking Water Consumed as such and as Tea (n=3073, 18 years and older)

Taste category

Good Not perceptible Weak, not object-

ionable Objectionable or

worse

Quantity of daily consumed tapwater in liter/person

Total

1.36 1.23

1.19

1.17

Water as such

0.32 0.20

0.20

0.15

Water as tea

0.32 0.31

0.30

0.31

The fact that water of which the taste and odour is disliked will be consumed in lesser quantities is obvious. However this human reaction has a certain function as it can contribute to the protection of the consumer against exposure to hazardous water contaminants. An additional point of interest is that several in-vestigations have reported a relation between the sensitivity of the chemical senses of a person and the needed dose of drugs to elicit a specific pharmacologi-cal effect (5,6). This means that the more sensitive part of the population, assessing water taste as objectionable, might also belong to the group most vulnerable to toxic effects. These considerations support the requirement that water should be tasteless, even for the more sensitive minority of the consumers. Of course these considerations do not imply that tasteless and odourless water will always be safe to drink as it may contain e.g. chemicals having carcinogenic effects after long term exposure, while the concentrations of these chemicals in water are not perceptible by the sense of smell.

Finally from the data presented in table 3 it is evident that consumers only avoid the consumption of the objectionably tasting water as long as this taste is per-ceptible. When it is masked, as is the case in coffee, soup or lemonade, the warning function of the chemical senses is no longer effective. As a result of this masking effect the influence of impaired water taste on the total quantity of daily consumed tapwater is relatively small. It seems to be worthwhile to study the eventual need to make people more aware of the warning function of the chemical senses in health protection.

SENSORY ASSESSMENT BY A PANEL OF 20 TYPES OF DRINKING WATER

In order to obtain a basis for comparison of sensorily perceptible quality aspects of certain types of drinking water and the presence of chemical compounds causing the observed differences, 20 public water supply systems were selected for further study. Among these supplies 8 used groundwater and 12 surface water as raw water source, while 7 supplies applied chlorination and 3 ozonisation. Half of the supplies used water directly or indirectly originating from the river Rhine. Water samples to be used for sensory evaluation were collected in 25-50 liter glass bottles and were presented during 3 sessions to a panel consisting of 52 subjects selected on the basis of a not too low odour sensitivity for aqueous solutions of isoborneol and o-dichlorobenzene. These subjects were as representative as possible with regard to the distribution of sex, age and domicile in relation to the popu-lation of The Netherlands (3). During each session 8 tapwater samples were presented to the panel 20 times. The subjects were instructed to smell and taste the sample, without swallowing it, and to note on a punch card the item on a taste

362 B. C. J. Zoeteman, G. J. Piet and C. F. H. Morra

(odour) scale which resembled most closely the perceived sensations. The des-criptive terms of the scale and the 1-dimensional solution for the psychophysical scale values are presented in table 4. These values were calculated by means of the method of triads (7) after treatment by multidimensional analysis (8).

TABLE 4 Items and Scale Values used for Taste and Odour Description by the Panel

Item number

1 2 3

4 5

Descriptive term

It tastes (smells) good It has no perceptible taste (smell) It has a weak, hardly objectionable taste (smell) It has an objectionable taste (smell) It tastes (smells) bad

Scale value

o.ooj 0.74 1.41

2.07 2.87

Also the character of the water taste and odour had to be noted by indicating which of the following items resembled most closely the perceived quality: chlorine like, earthy, putrid, metallic and faint unquantifiable. The preparation and realization of the panelsessionswas carried out in close co-operation with Prof. Dr. E.P. Köster and his collaborators of the Laboratory of Psychology, University of Utrecht, The Netherlands. As already shown by the results of the inquiry, the smell of the water was much more difficult to detect than the taste. Mean odour ratings varied only from 0.83-0.98 for the 1st session while mean taste ratings varied from 1.03-1.95 on the scale. A classification of the tested types of drinking water according to the mean taste rating by the panel is shown in table 5. This table confirms the data given in table 1 indicating a major role of the types of raw water on the taste of the final product. The average taste rating for ground-water supplies amounted to 1.17 and for surface water supplies to 1.70. Among the surface water supplies, those applying dune infiltration showed better taste ratings than the considered supplies based on storage reservoirs or bankfiltration. The data also show that bankfiltration of Rhine water generally results in drinking water with a bad taste, probably because bankfiltered water is not treated as contaminated surface water but as uncontaminated groundwater. A further striking finding in this respect is the high score for the "chlorine-like" taste quality for supplies using bankfiltration, while particularly in the cases of water type nr. 18 and nr. 20 no chlorination is applied during treatment. So the cause of this dominant taste note must be the presence of persistent chemicals originating from the Rhine river. It was the aim of the chemical analysis to identify particularly the chemicals responsible for these effects.

CHEMICAL COMPOUNDS DETECTED IN 20 TYPES OF DRINKING WATER

As already indicated by the conductivity figures in table 5, it is unlikely that inorganic constituents are important causes of impaired taste of the studied drinking waters derived from surface water. More detailed data, discussed elsewhere (1,4) suggested a possible role of hydrocarbonate- and magnesium-ions in taste differences between groundwater supplies. However the overall result confirmed the hypothesis that taste impairment of the studied types of drinking water must be mainly due to the presence of certain organic compounds.

To identify these compounds composite samples of tapwater distributed in 20 commu-nities were collected in specially designed 200 litres stainless steel vessels, allowing sampling without the risk of atmospheric pollutants being absorbed in the

Sensoriiy Perceptible Organic Pollutants 363

TABLE 5 Classification of 20 Tapwaters according to the Taste Rating by a Selected Panel and Data on Taste Qualification and Total Salt Content as measured by Conductivity

Tap-water type number

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15

1 16 17 18 19 20

Main raw water source

Ground water

+ + + + + +

+ +

Surface water

+

+ + + + + + + + + + +

1 Type of storage

Dune infiltr

Dune infiltr Dune infiltr Dune infiltr Dune infiltr Reservoir Bankfiltr. Reservoir Bankfiltr. Bankfiltr. Reservoir Bankfiltr.

Mean taste rating

1.12 1.16 1.16 1.16 1.19 1.19 1.20 1.21 1.23 1.30 1.48 1.55 1.59 1.68 1.70 1.81 1.91 2.01 2.06 2.06

% of panel members scoring in a taste quality category

Chlo-rine like

3 5 5 5 2 3 3 6 4 6 7 14 6 21 15 17 19 36 12 34

Earthy

13 13 15 12 15 11 15 8 12 14 30 21 21 15 20 28 14 9 27 17

Putrid

4 3 7 2 9 3 10 0 4 4 12 10 12 6 16 20 18 12 24 8

Metal-lic

12 21 17 12 20 20 15 17 8 15 18 17 26 14 21 17 30 25 18 19

Faint un-quali fia-ble

22 21 22 25 24 22 24 19 24 24 18 22 19 19 17 11 14 8 11 17

Con-ducti-vity (uS/ cm)

605 500 215 440 175 230 570 1140 190 875 ; 965 840 675 635 895 1065 590 805 790 860

sampled water. Samples for analysis of volatile organic compounds were collected separately in 1 litre glass bottles. The latter samples were tested for the pre-sence of lower halogenated compounds by injecting a head space sample on a glass capillary column, installed in a gasChromatograph with an Electron Capture Detec-tor (9) .

Water from the 200 litre vessels was used for a number of different determinations. Organochlorine pesticides were measured by means of G.C. and Electron Capture Detection, polynuclear aromatic hydrocarbons were analyzed by Thin-Layer Chromato-graphy and fluorescence detection (10,11). Organic compounds were concentrated from 5 litre samples by means of closed-loop gas stripping and carbon adsorption as developed by Grob e s . (12,13). The activated carbon filter was eluted with 25 ul of carbondisulphide and this concentrate was subsequently analyzed on a G.C.-M.S. system (Finnigan 3200) connected with an Interdata 70 computer (Wissen-schaftlichen Daten Verarbeitung, München). The G.C. system consisted of a 50 m OVioi glass capillary column and a splitless injection system and was coupled with the massspectrometer by means of a glass coated metal joint.

Furthermore the higher boiling and somewhat polar compounds were concentrated by adsorption on a 50-50 mixture of XAD-4 and XAD-8 resin (14). A quantity of 80 litres of the sample was processed using 25 ml of XAD-mixture in a glass colum (0=15mm) at a rate of 50 ml/minute.


The resin was eluted with ethylether and the extract was concentrated by a factor 20. Subsequently the concentrate was analyzed with the earlier described G.C.-M.S. system. After identification, quantitative estimates were based on seperate runs on an identical G.C. system and a comparison with the response for calibration mixtures of the same or closely related compounds.

In this way a total number of 280 compounds was detected of which however 94 could not be identified and 25 could only be tentatively identified. Although the majo-rity of compounds was detected in surface water supplies, it is striking that a number of halogenated compounds are present in groundwater supplies as shown in table 6. This table indicates that oxygen containing compounds are the largest group of detected substances (39% of the total number) followed by the hydrocar-bons (27%) and the halogen compounds (19%).

TABLE 6 Occurrence of Organic Compounds in 20 Types of Drinking Water

Type of Compound

Hydrocarbons bxygencompounds Halogencompounds Ni trogencompounds Sulphurcompounds Misee11.compounds

Total number

% of total number of identified compounds in a tapwater category

ground water sources

44 41.5 11 2.5 1 0

84

Surface water sources

bankfil-tration

35.5 30 21.5 9 0 4

79

dune in-filtration

36.5 29.5 21 7 2 4

101

Dpen re-servoirs

26.5 38 23 6 3 3.5

113

total

28.5 39 20 9 2 3.5

L65

all types

27 39 19 10 2 3

186

Furthermore table 6 suggests that the longer the surface water is exposed to light during treatment the more compounds, and particularly oxygen containing compounds, are formed. The high proportion of halogen compounds in water derived from open storage reservoirs is probably due to the coincidence that these supplies are also the supplies using break point chlorination during treatment. As table 7 shows hydrocarbons are most frequently detected in drinking water and particularly the lower substituted benzenes. Among the oxygen compounds dibutylphtalate is most frequently found and among the halogen compounds chloroform is most abundantly present. Chloroform is also the compound detected at the highest concentrations. As table 8 indicates this is due to application of breakpoint chlorination. The use of chlorination also results in the formation of iodine containing haloforms and in the production of compounds like 1,1 dichloroacetone, trichloroacetone and trichloronitromethane.

As taste problems are found most frequently for surface water supplies, it is of interest to consider those compounds detected at significantly higher levels in surface water supplies than in groundwater supplies. Such a comparison has been worked out in table 9. Of course not all compounds listed in this table are affec-ting water taste as the concentration of some of them will be far below the Odour Threshold or Taste Threshold Concentration. On the other hand taste relevant substances may not be listed in table 9 in case the estimated concentration in water derived from surface water was less than a factor 30 above the analytical detection level.


TABLE 7 Most Frequently detected Compounds in 20 Typesof Drinking Water

Type of compound

Hydrocarbons

pxygencompounds

Halogen compounds

Name

Toluene Xylenes C3-benzenes Decanes Ethylbenzene Fluoranthene Nonanes Naphtalene Bibutylphtalate 1,1 Dimethoxyisobutane Methylisobutanoate

ί Chloroform Tetrachloromethane

Detection frequency

20 1 19

19 18 17 16 15 14 17 13 13 16 15

Max. Cone. (ug/D

0.3 0.1 0.7 0.5 0.05 0.05 0.4 0.1 0.1 0.3 1.0

60 0.7

* General detection level: 0.005 ug/1

TABLE 8 Effect of Chlorination on Occurrence of some Halogenated Compounds in Tapwater

JType of chlorination

iNone Disinfection [Breakpoint chlorina-1 tion

Number of supplies studied

13 4 3

Concentration range in ug/1

Chloroform

(0.01-2.2 0.1 -10 25 -60

Dichloro-bromome-thane

<D.01-0.9 40.01-11 15 -55

Dichloro iodome-thane

40.01 CD.01-0.3 CD.01-1.0

Bromo form

dD.01

3..0-10

1,1 Di-chloro acetone

£.005 4).005 0.1-1.0

Tri~ chloro nitro methand

43.01 t).01-8.0 Q.0t-3.d

TABLE 9 Compounds detected in Drinking Water derived from Surface Water at Levels 30 or more Times higher than the Highest Levels found in Waters derived from Groundwater

compound name

Nonanes Octanols 1,1 Dimethoxypropane n-Hexylbutanoate Dichlorobromomethane Di chloroiodomethane Chlorodibromomethane Bromoform Chlorobromoiodome thane 1,2 Dichloropropane

Concentration ratio surface water supply ground water supply

40 300

> 60 > 200 550

> 100 1800 1000

> 30 > 30

Detection frequency in surface water supplies

8 6 3 1 9 4 9 7 4 2 1

366 B. C. J. Zoetman, G. J. Piet and C. F. H. Morra

TABLE 9 continued

Compound name

tp-Dichlorobenzene Trichlorobenzenes 1,1 Dichloroacetone bis-2-Chloroisopropylether Tetrachloro-di-isopropylether 2-Chloroaniline 5-Chloro-2-toluidine Trichloronitromethane Triethylphosphate

Concentration ratio surface water supply ground water supply

60 60 200 600 60 60 40 300 60

p-Nitrophenyl-2-chlorosulphone 60

Detection frequency] in surface water supplies

5 4 3 8 2 3 2 3 9 1

SELECTION OF TASTE IMPAIRING COMPOUNDS

As a first approach in selecting the most important taste impairing compounds the detected concentrations were compared with the individual Odour Threshold Concen-trations (OTC) as reported in literature (15). In this way additivity, senergy and masking effects are neglected. OTC values have to be used due to a nearly complete lack of taste threshold concentrations (TTC) for chemical compounds in water. As it is likely that the OTC will generally be lower than the TTC and because a.o. subthreshold levels are still relevant for the more sensitive minority of the population, concentrations above 1% of the OTC are considered to be of significance to taste. In this way a number of taste relevant compounds, possibly of biological origin were found, such as geosmin, octene-1 and several adehydes. A list of probable taste impairing compounds of industrial origin is given in table 10. The OTC data for the chlorinated anilines were measured afterwards at our laboratory as these compounds were particularly present in bankfiltered water. They indeed seem to contribute to the described chlorine-like taste of drinking water derived,from bankfiltrate.

TABLE 10 Industrial chemicals probable contributing to impaired taste of drinking water

Compound name

p-Dichlorobenzene Chloroform 1,3,5 Trimethylbenzene 2-Chloroaniline 1,2,4 Trichlorobenzene 5-Chloro-2-Toluidine y-Hexchlorocyclohexane Bromoform Hexachlorobutadiene 3,4 Dichloroaniline o-Dichlorobenzene bis-2-Chloroisopropylether

Odour Threshold oncentration(ug/1)

0.3 100 3 3 5 5 0.3

300 6 3 10

300

C /OTC max

1.0 0.6 0.2 0.1 0.06 0.04 0.03 0.03 0.02 0.01 0.01 0.01

i

!

1 * The OTC is the concentration of a compound in water at which the smell is still detectable for 50% of the observers.


Other important compounds, besides the chlorinated anilines and toluidines, are the chlorinated benzenes. The haloforms will not have a great effect on taste because of their rather sweet smell. The same applies to lower chlorinated com-pounds like tri- and tetrachloroethene, which have been detected recently in con-centrations in groundwater which closely approach the O.T.C.

As taste impairing compounds are part of the so-called grey list of chemicals of which the discharge into the surface water should be strictly controlled, according to the recently signed Convention against Chemical Pollution of the Rhine River (16), it is suggeste that the compounds listed in table 10 are incorporated in this grey list in order to reduce on short term their discharge into this river, which is so essential for the health and wellbeing of many millions of europeans.

REFERENCES

1 Zoeteman, B.C.J. (1978) Thesis on sensorily perceptible contamination in drinking water (in preparation)(University of Amsterdam)

2 Bartoshuk, L.M., NaCl thresholds in man: thresholds for water taste or NaCl taste?, J. Comparative and Physio!. Psychol., 87, 2, 310 (1974)

3 Zoeteman, B.C.J., Grunt, F.E. de, Koster, E.P., Smit, K.G.J., Punter, P.H., Taste assessment of salts in water, Chem. Senses and Flavor (in press)

4 Zoeteman, B.C.J., Piet, G.J., Morra, C.F.H., Grunt, F.E. de, Taste and chemical composition of drinking water, 6th International Symposium on Olfaction and taste, Paris, 15-17 July 1977, Information Retrieval Limited, London, (in press)

5 Joyce, C.R.B., Pan, L., Varonos, D.D., Taste sensitivity may be used to predict pharmacological effects, Life Sciences, 7, part I, 533 (1968)

6 Fischer, R., Gustatory, behavioral and pharmacological manifestations of chemo-reception in man, in Gustation and olfaction, edited by Ohloff, G. and Thomas, A.F., Academic Press, London, 187 (1971)

7 Torgerson, W.S., (1967), Theory and methods of scaling, 7th edition, Wiley, New York

8 Roskam, E., (1970), Nonmetric multidimensional scaling "minissa-l" version for triadic data (University of Nijmegen, Psychological Department)

9 Grunt, F.E. de, Piet, G.J., Keuvel, R. van de, Determination of volatile halogenated organic compounds in water by means of direct head-space analysis, Analytical Letters (in press)

10 Borneff, J., Kunte, H., Carcinogenic substances in water and soil, Part XXVI, Arch. Hyg. 153, 220 (1969)

11 Piet, G.J., Zoeteman, B.C.J., Klomp, R., Polynuclear aromatic hydrocarbons in the water environment of The Netherlands, RID-communication, 75-6(1975) (available at N.I.W.S., P.O. Box 150, 2260 AD Leidschendam, The Netherlands)

12 Grob, K., Grob, G., Grob, K.Jr., Organic substances in potable water and its precursors, Part III, J. of Chrom., 106, 299 (1975)

13 Grob, K., Zürcher, F., Stripping of trace organic substances from water, equipment and procedure, J. of Chrom., 117, 285 (1976)


14 Junk, G.A., Chriswell, CD., Chang, R.C., Kissinger, L.D., Richard, J.J., Fritz, J.S., Svec, H.J., Applications of resins for extracting organic com-pounds from water, Z. Anal. Chem., 282, 331 (1976)

15 Gemert,L.J. van, Nettenbreijer, A.H. (1977), Compilation of odour threshold values in air and water, (available at N.I.W.S., P.O. Box 150, 2260 AD Leid-schendam, The Netherlands)

16 Tractatenblad, nr. 31-34 (1977), Staatsuitgeverij, fs Gravenhage

Bioassays on Water Micropollutants

PHILIPPE LAZAR and DENIS HEMON

Unite de Recherches Statistiques de VInserm, Villejuif, France

ABSTRACT

Two methods of chloroformic extraction of water micropollutants are shown to give consistent information on the level of pollution of various kinds of waters· The biological effects of the extractible pool of chemicals are measured by two cytotoxicity tests and one in vivo test of cancer promoting activity· These bioassays have been applied to the study of the variations of micropollution along a river and to the comparison of flowing in and out of three drin-king water treatment stations·

INTRODUCTION

Water used for human consumption is contaminated by various organic micropollutants· The problem of studying their poten-tial health effects is by some aspects very similar to the study of the biological effects of cigarette smoke· Just ajs drinking water, cigarette smoke contains a great num-ber of chemicals among which some are known or suspected to be dan-gerous for health· However most of these substances, when identi-fied, are shown to be present in very small quantities, which raises the problem of their joint action and possible interaction in order to explain the overall biological activity of the smoke· In such conditions it is not easy to get a relevant information on the toxi-city of a given type of cigarettes· Two main ways have been used for this purpose : one of them is, in spite of the above-mentioned difficulties, chemical analysis of the smoke, searching for the most active families of substances (polycyclic hydrocarbons, nitrosami-nes, etc ···) ; the other way, which is more synthetic, is bioassay· Most of such assays use cigarette smoke condensate9 that is a pool of substances gathered from smoke by various methods· As anybody knows, the simple information on the quantity of tar produced by a cigarette - irrespective of its quality - is used for health pur-poses ^for instance it is printed, in some countries, on the ciga-rette packages)·

Coming back to water micropollution, it is of interest to try to ga-ther the whole - or a significant part - of the organic contaminants

369

370 P. Lazar and D. Hemon

of water in order first to know their total amount and also to try the measure their overall activity on some relevant biological sys-tems· It is clear that such an approach of health problems cannot be conducted without any chemical and epidemiologic validation· Ho-wever the aim of the present paper is to show that it can bring by itself some consistent information on the level and seriousness of water micropollution·

MATERIAL AND METHODS

Chloroformic extraction of water micropollutants is currently carried out in France since several years· Two methods have been principal-ly used : the first one is the extraction by chloroform in proportion 1 to lOat pH 7 ("simple extraction"), the second one has several steps more, using chloroform at pH 2 and 10 and different kinds of resins and charcoals ("total extraction") (l)·

Various kind of bioassays have been checked· Three of them will be reported here : two are in vitro cytotoxicity assays, the third one being and in vivo short term test of carcinogenicity· The first as-say uses cultures of a continuous human cell strain (Hela cells) (2), the second one uses cultures of hamster fibroblasts (3)· The short term tests of carcinogenicity are those which are currently used for the detection of promoting activity of chemicals or of ci-garette smoke condensates (4> 5) * they are based on the destruction of sebaceous glands of mouse skin and its hyperplasia (6)·

For each cytotoxicity test it is possible to get an efficient dose 50 % (ED 50 %') in its conditions of utilization, i· e. the dose which inhibits the growth of half the cell clones· For the mouse skin tests, the results are expressed in percentages of decrease of the number of sebaceous glands and of increase of the thickness of epidermis, in comparison with controls· Some of the results are ex-pressed in relative potencies, i· e· ratios of efficient doses·

RESULTS

Three main studies will be reported· The first one was organized to compare 40 different waters from various origins· It gives an infor-mation on the distributions of the measures according to these ori-gins· The second study allows to compare the evolution of water mi-cropollution along a french river· The third one gives a glance on the problems which are raised by the treatment of superficial water in order to convert it into drinking water·

Amount and Cytotoxicity of Micropollutants of 40 Kinds of Waters

Table 1 gives the amounts of micropollutants detected by both methods of extraction, applied to aliquot samples· Drinking waters carry significantly (ρ < 0.001) lower amounts of micropol-lutants (about four to five times less in simple extraction, twice to three times less in total extraction)· Figure 1 shows the corre-lation between both measures (r = 0.69* P ^ 0#00l)·

Table 2 gives the intrinsic cytotoxicity of the micropol-luants extracted from these 40 waters, measured by the first above-mentioned biossay (Hela test)· The ED 50 % is significantly lower

Bioassays on Water Micropollutants 371

TABLE 1 - AMOUNTS**' OF MICROPOLLUTANTS IN 10 WATERS

SUPERFICIAL WATERS

DRINKING WATERS

TREATED SUPERFICIAL WATERS

DEEP WATERS

n

24

11

5

SIMPLE EXTRACTION

0.66 t 0.07

0.16 ± 0.02

0.10 + 0.05

TOTAL 1 EXTRACTION

2.29+0.23

1.22+0.19

0.77 t 0.14

( * ) - MEAN VALUE ± STANDARD ERROR (MG/L)

mg/l

o "■♦-o σ

I

* · · ·

£■ *+

^· ·

me/I

Simple extraction

Fig. I.

Correlation between amounts of mlcropollutanta from aimple and total extraction

( · superficial water, * drinking water)

372 P. Lazar and D« Hemon

for superficial waters (p < 0#001 for simple extraction), which means what drinking water simple extracts have, in the average, a smaller cytotoxicity in Hela test than those from superficial waters· The differences are not significant for total extracts· Figure 2 shows the correlation of the cytotoxicities of the two kinds of extracts <r = 0,37, P - 0·05)·

ED 50

S

600

500

400

300

200

100

0

mg/|

•

•

t *

•

•

* * •

•

• •

•

*

•

*

*

* *

•

1

*

1

*

100 200 300 400

Simple extraction

Fig. 2. Correlation of the cytotoxicities of simple and total extracts

( · superficial water, * drinking water)

mg/|

ED 5,0

Bioassays on Water Micropol lu tan ts 373

TABLE 2 - INTRINSIC CYT0T0XICITY(* ) OF MICROPOLLUTANTS FROM 40 WATERS

(HELA TEST)

SUPERFICIAL WATERS

DRINKING WATERS

TREATED SUPERFICIAL

WATERS

DEEP WATERS

n

24

11

5

SIMPLE EXTRACTION

115 ± 14

225 ± 44

262+ 23

TOTAL EXTRACTION |

172 + 25

233 ± 32

240 i 46

( * ) - EFFICIENT DOSE 50 % ± STANDARD ERROR (MG/L)

Figure 3 shows that there is no correlation between the amount of total micropollutants and their intrinsic activity (r = 0·15, NS) · However Pig· 4 shows that micropollutants from simple extraction have a negative correlation between their amount and their ED 50 % (r = - 0·34, ρ = 0 # 05) i. e· a positive correlation between amount and cytotoxicity on Hela test· But the correlation results essen-tially from the fact that drinking waters are less rich in micropol-lutants then superficial waters (Table 1 ) , and that these micropol-lutants are, in the average, intrinsically less active (Table 2 ) . Within each category of waters, there is no residual correlation (r = 0.12 and r = 0·ΐ6 respectively, NS) ·

Evolution of water micropollution along a french river (simple ex-traction only)

Figure 5 shows the variations along a river of the amounts of micro-pollutants and of their biological activities (measured by the three above-mentioned bioassays)· Eight points of sampling have been used (including one point before and one after a station of water treat-ment) · Values are given in relative potencies by comparison with the first point of sampling (the nearest of the spring)·

Effects of treatment of superficial water on the amount of micropol-lutants and their activity (total extraction only)

Figure 6 shows the quantitative reduction of total micropollution in the three stations of treatment of superficial water for its conver-sion into drinking water. In the average the reduction is about one half, but it is not as high when the initial level of pollution is rather low· As far as biological activity is concerned, the three bioassays do not give similar results : the Hela test shows, as a general trend, a lower intrinsic activity for micropollutants from treated waters (Table 3 ) , when the two other tests show a reverse trend (Tables 4 and 5)·

374 P. Lazar and D, Hemon

6 0 0

500

4 0 0

t5 e 8 300

,o

2 0 0

100

ED 50 mg/l

* ·

»· ·

5mg/ l

Simple extraction

Fig. 3. Correlation between the amount of total mlcropollutants and

their intrxnsic activity

( · superficial water, + drinking water)


500

400

■$. 300

200

1001

ED 50 mg/|

·· ··· • ·

0-5 1-0 1.5

Simple extraction

2-0 mg/l

Fig. 4. Correlation between the amount of micropollutants from simple extraction

and their intrinsic activity

( · superficial water, + drinking water)

P. Lazar and D. Hemon

Amount of micropollutants

57/^^^^-;^*.''υ^*ϊ-ύ:'./*-·»; '..*?**.*";,:<■■.:,£■:

Heia cytoxicity test

Fibroblast toxicity test

•;v Jfy;V·-;·· '

l i fS»*«!i

Mouse hyperplasia test

Mouse sebaceous gland test

Lock Flow of the river - * - Water treatment station

Fig. 5

Micropollution along a french river

376

Bioassays on Water Micropol lu tan ts 377

4 AMOUNT ( m g / l )

100 J

0.0

*## * # # NS

Legend : F l o w i n g i n

F l o w i n g o u t

S i g n i f i c a n c e o f t h e c o m p a r i s o n b e t w e e n f l o w i n g i n and o u t : s e e T a b l e 3

F i g . 6

Mean amount of micropollutants (total extraction) in 3 treatment stations (flowing in and out)


TABLE 3 - CYTOTOXICITY TESTS ON HELA CELLS OF TOTAL EXTRACTS FROM FLOWING IN AND OUT WATERS IN 3 TREATMENTS STATIONS (EFFICIENT DOSE 50 % ± STANDARD ERROR)

FLOWING IN

FLOWING OUT

LEVEL OF SIGNIFI-CANCE BETWEEN FLOWING IN AND OUT

STATION A

260 + 37

329 + 11

NS

STATION B

215 ± 33

337 ± 29

* *

STATION C

313 ± 38

370+ 37

NS

LEVEL OF SIGNIFICANCE 1 BETWEEN THE 1 3 STATIONS |

NS

NS

LEGEND : NS ¥ ¥¥ ¥ ¥¥

NON SIGNIFICANT P - 5 %

P - 1 %

P £ 1 "/··

TABLE 1 - CYTOTOXICITY TESTS ON HAMSTER FIBROBLASTS OF TOTAL EXTRACTS FROM FLOWING IN AND OUT WATERS IN 3 TREATMENT STATIONS (EFFICIENT DOSE 50 % ± STANDARD ERROR)

FLOWING IN

FLOWING OUT

1 LEVEL OF SIGNIFI-CANCE BETWEEN

FLOWING IN AND OUT

STATION A

15 + 15

13 + 3

•

STATION B

33 + 9

11 ± 3

•

STATION C

1 2 + 3

12 ± 3

NS

LEVEL OF SIGNIFICANCE 1 BETWEEN THE 3 STATIONS

• •

NS

LEGEND : NS ¥ ¥ ¥

¥ ¥ ¥

NON SIGNIFICANT Ρ = 5 ί p = 1 %

p 1 V··


TABLE 5 - SEBACEOUS GLAND TEST AND HYPERPLASIA TESTS OF TOTAL EXTRACTS FROM FLOWING IN AND OUT WATERS IN 1 TREATMENT STATION

(EFFICIENT DOSE 50 Z ± STANDARD ERROR)

FLOWING IN

FLOWING OUT

1 LEVEL OF SIGNIFI-CANCE BETWEEN

FLOWING IN AND OUT

SEBACEOUS DECREASE OF THE NUMBER IN COMPA-

RISON WITH CONTROLS (%) +STANDARD ERROR

40.il ± 10.4

99.4 + 0.3

+ ¥ ¥

GLAND

LEVEL OF SIGNIFICANCE

* * *

* ¥ *

HYPERPLASIA 1 INCREASE OF SKIN | THICKNESS IN COM- LEVEL OF PARISON WITH SIGNIFICANCE CONTROLS (%)

+STANDARD ERROR | |

113.4 + 18.8

597.1 + 49.9

¥ ¥ ¥

* ¥ ¥

¥ ¥ ¥

LEGEND : NS ¥

¥ ¥

¥ ¥ ¥

DISCUS

NON SIGNIFICANT P - 5 % P - 1 % p 1 "/··

SIQN

As Fig· 1 and Fig· 2 show, there is a rather good correlation between the two ways of measuring water micropollution, by simple or by to-tal extraction· The amounts of micropollutants they gather and their activities are not at the same level, but both can be used in order to compare similarly different kinds of waters· Intrinsic cytotoxici-ty (on Hela cells) is slightly smaller with total extracts, which im-plies that some of the components added by comparison with simple extraction are less active· This assessment can be verified directly by studying cytotoxicity of the main components of total extract, i· e· itsthree constitutive fractions : neutral, acid and basic· Ta-ble 6 shows that the acid fraction in the average is much less active that the neutral and basic ones (p < O.OOl)·

Total extraction, which looks to be quite more efficient a priori, is then not necessarily much better than simple extraction to get information on the global level of activity of the micropollutants of a given water (at least on the Hela test)· It must be noticed al-so, from fig· 3 and 4, that there is no correlation between the quan-tity of micropollutants and their intrinsic activity at least within a class of waters (superficial waters, drinking waters)· Both infor-mations are then necessary in order to assess the toxicity of any water·

Also from this point of view, Fig. 4 shows that the variations of mi-cropollution along a river are susceptible to be measured, and to show variations from one place to another· For instance it is of in-terest to notice that there is a clearcut accumulation of micropol-lutants at the level of the lock which is located on the figure, but


TABLE 6 - MEAN PROPORTION AND CYTOTOXICITY OF THE VARIOUS EXTRACTED FRACTIONS (HELA TEST)

PROPORTiON( * '

IN COMPARISON WITH TOTAL

EXTRACT

EFFICIENT D O S E ( * ' 50 % (MG/L)

SIMPLE EXTRACT

m % + 6 %

160 + 18

1 - NEUTRAL | FRACTION

33 % + 7 X

m + lo

2 - ACID FRACTION

53 % + 11 %

323 ± 26

3 - BASIC FRACTION

11 % + 6 %

189 ± 18

MOTAL EXTRACT Ί ( 1 + 2 + 3 )

100 %

196 + 19

(*) + STANDARD ERROR

that these supplementary micropollutants are not very active on all the biological systems used : some intrinsic activities are even lo-wer at this place· Hovewer, generally speaking, toxicity increases regularly along the river, qualitatively and quantitatively· The last point, just at the flowing out from a water treatment station, is of peculiar interest· The amount of micropollutants is greatly decreased, but their intrinsic activity is significantly higher than at the flowing in·

The last observation can be related to questions which are now raised about the influence of water treatment on the biological properties of its components and especially its residual micropollutants· When it is clear that, in the average, water treatments reduce the total amount of micropollutants {Fig· 6 ) , it is not obvious that they should also reduce their biological activity, since they have not been planned for this purpose· And, in fact, Tables 3, 4 and 5 show that some activities are reduced, some other enhanced· Further stu-dies are then necessary in order to determine whether some water treatJ-ments should be modified· Chemical analysis can obviously be of great help in such studies, but does not suppress the possible contribution of bioassays, such as those which have been described, as guidelines for the improvement of the quality of drinking water·


REFERENCES

(1) R· Cabridenc, A· Sdika, Quelques aspects de lfextraction et de 1*identification des micropolluants des eaux· Techniques et Sciences Municipales 7, 285 (1975)

{2) C· Gerin-Roze, D· Homon, C· Berger, B. Festy, Utilisation de cultures de cellules de lignee continue (Hela) pour la mise en evidence de potentialites toxiques in vitro, En preparation (1978)

(3) I· Chouroulinkov, C. Lasne, Cancerogenese chimique en culture de tissus : criteres et tests de transformation· Serie Sympo-sia INSERM 52, 207 U976)

(4) 1· Chouroulinkov, P· Lazar, C· Izard, C. Liberman, M· Guerin, Sebaceous gland and hyperplasia tests as screening methods for tobacco tar carcinogenesis· J· Nat« Cancer Inst· 42, 981 U969)

(5) P· Lazar, I· Chouroulinkov, Validity of the sebaceous gland and the hyperplasia test for the prediction of the carcinogenicity of cigarette smoke condensate and their fractions· 0974) Experimental lung cancer KARBE and PARK, Springer Verlag·

^6) Ρ· Lazar, C· Liberman, I· Chouroulinkov, M· Guorin, Tests sur la peau de souris pour la determination des activites carcinoge-nes, Bull» Ass· FranQ· Cancer 50, 567 (1963)

■it

This work has been achieved by a group of laboratories including : Institut de Recherches Chimiques Appliquees (R· CABRIDENC, A· SDIKA), Laboratoire d'Hygiene de la Ville de Paris (B. FESTY, C. GERIN-ROZE), Laboratoire de Medecine Experimentale du CNRS (l# CHOUROULINKOV), Unite de Recherches Statistiques de 1*INSERM t'C. BERGER, D. HEMON, P. LAZAR), Societe d*Etudes pour le Traitement et 1*Utilisation des Eaux (C. GOMELLA), Societe Lyonnaise des Eaux et de l*Eclairage <J.J· PROMPSY) with grants from the Ministere de la Culture et de 1*Environneraent (A· YANA)·

Research in the United Kingdom on Health Aspects of Wastezvater Reuse for Potable Supply

R. F. PACKHAM

BSc, PhD, CChem, FRIC, FIWES, MInstWPC

ABSTRACT

In the United Kingdom, approximately one third of potable water supplies are derived from lowland rivers used also for conveying treated domestic and industrial wastes to the sea. These rivers are relatively short and often support large populations and considerable industrial activity. The proportion of the total river flow due to wastewater can therefore be high, particularly during periods of low river flow.

In anticipation of probable further increase in the extent of this indirect reuse of wastewater for potable supply, investigations have been commissioned by the Department of the Environment into possible health effects of trace organic substances in drinking water. The research programme includes analysis of water supplies to determine the nature of any compounds present in drinking water as a direct result of contamination of the source with wastewater. Available mutagenic screening tests will be adapted and applied to the examination of water supplies for the presence of mutagens and identification of substances responsible for any positive responses. The incidence of certain chronic diseases, particularly cancer of the digestive organs and genito-urinary system, will be examined in relation to the extent of wastewater reuse for potable purposes in different towns in the UK.

INTRODUCTION

Nearly a century ago dramatic improvements in public health followed recognition of the importance of protecting public water supplies from human waste. During the past 50 years, however, the practice of drawing water supplies from sources contaminated to some extent with sewage effluent has increased in the United Kingdom. The depletion of groundwater sources and the high cost of conveying unpolluted upland waters over great distances has led in some cases to a reassessment of water sources previously considered doubtful or even unaccept-able for public water supply. The growth of the population and industry in lowland areas has also resulted in an increase in the level of contamination of existing lowland sources of water with sewage and industrial effluent. The main potential health hazard of such a situation is undoubtedly contamination of water supply with bacterial pathogens, but public health records have amply confirmed

383

384 R. F. Packham

the efficacy of modern disinfection technology in these s i tuations.

An important distinction should be made between the situation in the United Kingdom and that in some other countries where s c h e m e s have been or are being implemented to produce drinking water by the direct p r o c e s s i n g of sewage effluent. Such s c h e m e s (direct wastewater reuse ) invariably involve advanced mul t i - s tage treatment sequences in contrast to the indirect reuse situation in the United Kingdom where in many c a s e s a s lowly developing i n c r e a s e in pollution l e v e l s affects sources subject to conventional water treatment .

Increased awarenes s of health effects a s s o c i a t e d with environmental contaminants has l ed to concern that the recyc l ing of effluent wil l resul t in a build up of certa in refractory substances , some of which could be of health s ignif icance. The potential problems have been rev iewed by a number of groups, including an international working meet ing sponsored by the World Health Organisation International Reference Centre for Community Water Supply (1). The report of this meeting set out proposals and a s s i g n e d pr ior i t i e s for r e s e a r c h into health effects relating to direct and indirect reuse of wastewater for human consumption.

The increas ing tendency in the UK to use water s o u r c e s , contaminated to some extent with industrial and domest ic wastewater , has l ed to proposa ls for r e s e a r c h to determine the nature of drinking water contaminants derived from effluents and their health ef fects . R e s e a r c h p r o g r a m m e s are now being implemented to deal with these problems and the main purpose of the present paper i s to descr ibe this work. Some background information i s n e c e s s a r y , however , on the present l eve l of reuse in the United Kingdom.

Reuse in the United Kingdom

Sources of water used for public supply in the United Kingdom are , in general , of three kinds: groundwater s o u r c e s , unpolluted surface sources (general ly in upland moorland areas ) and reuse r i v e r s (usually lowland r i v e r s ) . A s a part of a rev iew of the extent of present r e u s e , the Central Water Planning Unit are preparing a l i s t of all water abstract ions for potable supplies from reuse r i v e r s in England and Wales together with their effluent content at average r iver flow. Further details include the amount of water that was actual ly abstracted in 1975 and the amount that can be taken at present by l i c ence . Tentative e s t i m a t e s of l i c e n s e d and actual abstract ions from the three different types of source are given in Table 1.

Research in the United Kingdom 385

TABLE 1 Sources of Water in England and Wales (expressed as millions of cubic metres per annum)

Unpolluted Surface Reuse

Abstraction Groundwater Water Rivers Total

Licensed amount (1976-7) 2925 3332 3273 9530

Actual abstraction (1975) 1852 2201 1557 5610

There has been a threefold increase in the volume of water taken for drinking purposes from reuse rivers since the early 1930s. The volume of water abstracted from such sources in 1975, 1557 million cubic metres, contained on average 11% of industrial and domestic effluent, i. e. , 171 million cubic metres approximating to 5% of the total resource abstracted from surface and ground-waters in England and Wales. On some rivers the proportion of effluent can approximate to the total river flow in dry periods, but details of the proportion of effluent abstracted at such times are difficult to calculate due to operational variations. The most representative figures for the proportion of effluent in sources of water for public supply is given by the proportion at average daily flow. Examples of the proportion of effluent at some important points of abstraction are 40% in the River Lee at Chingford, 20% in the River Thames at Chertsey and 13% in the River Great Ouse at Offord. The proportion of all the reuse river abstractions in 1975 taken from rivers of different mean effluent content is shown as a cumulative curve in Fig. 1. In 1975, none of the rivers abstracted contained more than 41% effluent at mean flow while 35% of the total volume abstracted contained less than 5% effluent.

Although information on the quantity of effluent in water supply sources is of great importance it does not in itself provide information on actual chemical pollution levels. Further work is being undertaken by the Central Water Planning Unit to classify the types of effluent present and to determine the chemical quality of reuse river waters before and after treatment. Such information will be very useful but it will not enable a rigorous assessment of potential health problems to be made at this stage.

Water supplies in the United Kingdom, particularly in England, are dependent at the present time, therefore, on the use of many sources contaminated to some degree with sewage and industrial effluent. There is ample evidence that the indirect reuse of waste water is likely to increase both through increased levels of contamination of existing water supply rivers and possibly also through the use of sources that have not so far been regarded as suitable. The latter present a special problem common to other new materials (e.g. , drugs, water supply treat-ment chemicals and constructional materials) requiring clearance from the health aspect. This is that in general a much higher level of safety assurance is required to secure approval of a new material than was necessary for an existing material of a similar but not identical type. The same is true of sources of drinking water.

386 R. F. Packham

100-1

ω z O o < cc \-co CO < LU CO Ό dl CL

90-

8 0 J

_ 70- | Q)

E ■§ 6 0 ->

£ 5 0 J

o ig 4 0 -

30J

cc O a. O GC QL

20-

10-

"T"

5 —r~ 15

— i —

2 0 — i —

25 ~40~ 10 15 2 0 25 3 0 35

TOTAL EFFLUENT CONTENT AT AVERAGE FLOW (%)

— i 45

Fig. 1. Proportion by volume of reuse abstraction in UK (1975) related to river effluent content

In the UK there has in the past tended to be a sharp division between heavily polluted rivers and those rivers used for public water supply which have been to a large extent protected from substantial contamination. The introduction in recent years of stricter pollution control measures with consequent improvements in river water quality has caused some Water Authorities to look carefully at some rivers previously regarded as unacceptable for public supply. The River Trent, a major river not used for public supply, passing through the industrial Midlands of England,represents a substantial potential water resource in an area where this is much needed. A comprehensive research programme, part funded by the DOE, was undertaken by the Water Research Association and the Trent River Authority into the feasibility and costs of treating River Trent water to potable standards. Although this work showed (2) that,using a series of unit processes in addition to conventional treatment, raw River Trent water could be purified to meet most


accepted criteria for potable water, it was not possible to state that the water so produced was Wholesome1, this being the legal requirement for a potable water in the UK.

We have here a clear example of the failure of conventional criteria to provide an acceptable basis for the assessment of a new drinking water source. The need for information on health risks which could provide a basis for evaluating such sources in the UK is highlighted by the following statement in a report on Water Resources in England and Wales (3).

!Several important rivers are not at present suitable for providing water for public supply because of the volume and nature of the effluent discharge to them. If the quality of such rivers were improved sufficiently, they could make very substantial additions to water resources. The Trent is an outstanding example of a polluted river which could make a contribution in this way. T

Organic Compounds in Drinking Water

Concern at the lack of information on the nature of organic micropollutants in sewage effluents and lowland rivers (4) prompted the initiation of research programmes in the early 1970s at the Water Pollution Research Laboratory and the Water Research Association* sponsored by the Department of the Environment. At both laboratories the emphasis of much of the early work was on the development of analytical methodology for trace organic compounds. The techniques have since been applied to a variety of waters with a much greater emphasis recently on drinking water quality.

Considerable work was undertaken on isolation, concentration and pre-separation techniques for trace organic compounds in water and effluent samples. Methods used for identification and measurement included gas liquid chromatography (including the use of high resolution capillary columns), high performance liquid chromatography and gas chromatography coupled to mass spectrometry. Currently, special emphasis is being given to techniques for the analysis of non-volatile organic compounds which form a large part of the organic content of drinking water and for which good analytical techniques are lacking. In much of this work the Water Research Centre has been closely involved in the COST 64b

^Following reorganisation in 1974 these laboratories became the Stevenage and Medmenham laboratories, respectively, of the Water Research Centre.

388 R. ¥. Packham

Micro pollutants Projec t which involved col laboration and information exchange between 29 laborator ies in 12 countries on fthe development of methods for analysing as complete ly as poss ib le the organic pollutants present in a sample of water1 .

A s an important part of the project , the Centre has produced a comprehens ive l i s t of approximately 1600 organic compounds so far identif ied in waters of various kinds including drinking water (5). The l i s t , which has now been put on a computer base to faci l i tate a c c e s s , c o v e r s world l i terature and includes for each compound W i s s w e s e r - L i n e Notation code, sample detai l s , date of sampling, method of ana lys i s , concentration found and information source . Currently the l i s t includes over 350 compounds that have been identified in drinking water in different countr ies .

A number of important conclus ions can be drawn from an examination of this l i s t .

1 . A very wide range of different types of organic compound have been found in drinking water - there are entr ies under 29 different c l a s s e s of compound.

2. The concentration of the substances found i s very low, commonly l e s s than one part in a thousand mi l l ion .

3. There i s an overwhelming preponderance of volat i le organic substances amenable to analys i s by techniques based on gas chromatography. This fact re f lec ts the state of the art on analytical methodology rather than the kind of substances present in water . Other evidence indicates that volat i le substances represent only 1 0 to 20% of the compounds present .

An important considerat ion in relat ion to the a s s e s s m e n t of health hazards a s s o c i a t e d with the p r e s e n c e of organic compounds in drinking water i s the availabil ity of relevant toxicological data. Although a rev iew of the current l i s t of compounds found in drinking water from this aspect has not yet been completed, examination of an ear l i er l i s t (1) revea led a paucity and in some c a s e s a total lack of toxicological data. Part icular ly relevant was the lack of data on chronic toxicity, carcinogenic i ty , mutagenicity and teratogenic i ty and on toxic effects in general a s s oc ia t ed with very low concentrations of obnoxious m a t e r i a l s .

In the UK it has been recognised , therefore , that detai led analytical data for drinking waters wil l not by i t se l f enable an a s s e s s m e n t to be made at the present t ime of r i sks to health a s s o c i a t e d with organic compounds. The most general problem i s the shortage of toxicological data but of great consequence a l s o i s the inability of current analytical techniques to deal with the greater part of the organic matter in water. Although the analytical approach i s of great importance , a val id a s s e s s m e n t of poss ib le r i sks to health from organic compounds in water requires other work involving toxicological and epidemiological p r o g r a m m e s . These a spec t s have therefore been included in the UK invest igat ions into health a s p e c t s of wastewater reuse .


Toxicology

The purpose of a toxicological component in the research programme was to cover compounds not so far identified in water or identified compounds of unknown toxicology. To meet this requirement an overall assessment of the toxicity of the total mixture of drinking water components was required.

Detailed consideration was given initially to the possibility of using concentrates of water components in long-term animal feeding tests . Such tests, which would need to have extended over two or three years, would have necessarily involved very large numbers of animals in order to obtain statistically valid results. Apart from the astronomical costs which would have severely limited the number of supplies that could be studied, there were considerable technical problems involved in such a programme. Very large quantities of concentrate would have been required and it would have been difficult for several reasons to have avoided the possibility of changes in the composition of the concentrates being fed over the period of test. There were significant problems also in relation to the method used for concentrate preparation. This approach was therefore abandoned, an important additional factor in this decision being increasing confidence in the value of short-term mutagenic screening tests as a means of predicting carcinogenicity.

A variety of short-term screening tests for carcinogens have been described (6) and the Ames test in particular has been evaluated in some detail (7)(8). Current data provides evidence that carcinogens may be predicted on the basis of such tests with 90-93% confidence in contrast to a predictability of about 80% of long-term feeding tests based on a single mammalian species. A fluctuation assay procedure developed at the MRC Cell Mutation Unit (9) appeared to be particularly suitable for detecting carcinogens in water. This procedure, which involves measuring mutation to tryptophan independence in Escherichia coli in an aqueous medium, can incorporate metabolic activation systems and for direct acting mutagens a sensitivity of up to 100 times that of conventional tests is claimed. The procedure is simple and should be capable of further development to improve the sensitivity further and to automate many of the operations involved. A particularly attractive feature of the fluctuation assay procedure is that it is possible, in view of its sensitivity, that it can be applied directly to unconcentrated water samples thereby eliminating the uncertainties involved in the use of concentrates.

As a part of the UK research programme on health aspects of wastewater reuse, the MRC Cell Mutation Unit are investigating the adaptation of their mutagenic screening procedure for the examination of water samples. Following suitable refinement of the methodology it will then be applied to a range of water samples as a pilot to a more extensive survey of water supplies for the presence of mutagens. This survey will be designed to take account of type of water source with emphasis on extent of wastewater reuse (but also including some upland and groundwater sources), seasonal variations in water quality and water quality changes within a distribution system.

390 R. F. Packham

Mutagenic screening t e s t s wil l be undertaken on water samples from s e l e c t e d supplies and any posi t ive re su l t s wil l tr igger studies a imed at determining as far as poss ib le the chemica l nature of the mutagenic agent. This work wil l involve sys temat ic fractionation of the original water sample , further mutagenicity test ing and detai led chemica l ana lys i s of act ive fract ions . Work on the development and refinement of suitable sample fractionation procedures including adsorption onto XAD r e s i n s and separation of volat i le components by stripping techniques i s currently in p r o g r e s s . Identification of components wil l depend heavi ly on gas chromatography-mass spectrometry-data p r o c e s s i n g and poss ib ly field desorpt ion-mass spec trometry .

The outcome of this work should be:

(a) an a s s e s s m e n t of the extent to which potentially carc inogenic substances are present in drinking water;

(b) an a s s e s s m e n t of the relat ionship, if any, between the presence of c a r c i n o -genic substances in drinking water , the degree of wastewater r e u s e and other water quality re lated fac tors .

(c) a l i s t of carc inogenic and potentially carcinogenic substances found in drinking water .

The use of the term'potent ial ly carc inogenic ' impl ie s that in the course of this work mutagenic activity may be a s s o c i a t e d with compounds not recogn i sed prev ious ly as mutagens or carc inogens . The importance i s r e c o g n i s e d of undertaking further studies at a later stage on such compounds to conf irm carc inogenic propert ies and to evaluate d o s e - r e s p o n s e re lat ionships . Such work would include the use of additional screening s y s t e m s using bacterial and mammal ian c e l l s poss ib ly leading on to mammal ian carc inogenic i ty s tudies .

Epidemiology

In the study of wastewater reuse for potable purposes it i s c l ear ly of importance to determine whether there i s any evidence of regional variat ions in d i s e a s e incidence a s s o c i a t e d with the l eve l of source contamination. In the United Kingdom a comprehens ive drinking water quality data base for towns and certa in town aggregates having a population greater than 50 000 has been developed by the Water R e s e a r c h Centre as a part of an invest igat ion into the re lat ionship between water quality and cardiovascular d i s e a s e being undertaken in col laboration with the Department of Clinical Epidemiology and Social Medicine, Royal F r e e Hospital , London. This data base i s being extended to include certa in raw water quality data and information on r iver and effluent flows co l l ec ted from Water Authori t ies and other organisat ions including the Central Water Planning Unit. This information wil l be used to formulate indices of source contamination which wi l l be examined in relat ion to the incidence of certa in d i s e a s e s .


The reuse indices will include effluent flow as a proportion of river flow, contributing population upstream of the abstraction point and a series of chemical parameters. The epidemiological studies, which are being undertaken at the Royal Free Hospital, are concentrating on cancer of the digestive organs and peritoneum, the genito-urinary organs and the bladder but other possible disease outcomes will be investigated to provide a basis for the formulation of further hypotheses.

Because of the long induction period of diseases such as cancer, this type of study requires historical water quality data which is often difficult to obtain. The current water quality data base extends back to 1969 and it is hoped to extend this to cover at least a ten-year period for a number of situations. Apart from London there are sixteen towns which have been reusing water since before 1926 to some extent. Nine towns started reusing water between 1926 and 1935, seven more between 1936 and 1945 and eight between 1946 and 1955, twenty-three between 1956 and 1965 and nine others more recently. It is important to recognise that, as shown in Fig. l ,the actual level of source contamination in the majority of these situations is very low indeed. It follows from this and from the fairly recent introduction of some major reuse abstractions that the absence of any association between reuse and disease incidence at the present time will not provide conclusive evidence as to the absence of health effects and that further studies will be necessary in the future.

Overall Research Programme

The investigations described all form part of a four-year programme of research co-ordinated by the Water Research Centre and funded by the Department of the Environment. The work comprises the first two phases of a three-phase programme, the time scale for which is shown as a bar chart in Fig. 2.

The first phase consists of preliminary epidemiological work, refinement of mutagenic screening procedures and techniques of sample fractionation and the preparation of detailed plans for Phase 2. These will incorporate further epidemiological work and a survey to determine the level of contamination of water supplies with mutagenic substances and the nature of any such substances. Provision has been made throughout the programme for water quality studies in which advanced analytical techniques will be applied to survey the types of organic compounds present in drinking water as a result of contamination of the source with sewage effluent and industrial waste. These will be supported by detailed studies of the composition of specific effluents affecting water supply rivers.

It is anticipated that Phase 2 will lead to a requirement for follow up work in Phase 3, the nature of which cannot be defined at present but which would certainly include any confirmatory toxicological work as already mentioned.

If the work brings to light any potentially hazardous situations, there are a number of possible courses of action that may be required including changes in

392 R. F. Packham

water and effluent treatment p r o c e s s e s , the introduction of t ighter pollution control m e a s u r e s or even a drast ic change in the type of source in u s e . The Central Water Planning Unit i s at present considering the impl icat ions of poss ib le health hazards a s s o c i a t e d with water supply abstract ions from reuse r i v e r s and i s examining their potential effect on national s t ra teg ie s for water r e s o u r c e development.

Phase 1

Epidemiology - pre l iminary studies

Toxicology - adaptation of screening tes t

Refinement of concentration and fractionation techniques

Micropollutant studies in reuse situations

Detai led planning for Phase 2

Phase 2

Epidemiology Mutagenie screening of water

supplie s Identification of mutagens Micropollutant studies in

reuse situations

Phase 3

Further toxicology on substances identified in survey

Other follow up studies

1976 1977 1978 1979 1980 1981

F ig . 2. Bar chart for invest igat ion into health a spec t s of wastewater reuse in UK

In anticipation of a poss ib le need for improved methods for the removal of organic compounds from drinking water, potentially suitable p r o c e s s e s are being studied in detail at the Water R e s e a r c h Centre with particular emphas i s on the use of act ivated carbon.


ACKNOWLEDGEMENTS

The r e s e a r c h programme descr ibed in this paper i s being funded by the Department of the Environment. Information on the present extent of reuse in the United Kingdom was provided by the Central Water Planning Unit and i s gratefully acknowledged. The author w i s h e s to thank the Director of the Water R e s e a r c h Centre for p e r m i s s i o n to publish this paper.

REFERENCES

1. World Health Organisation, International Reference Centre for Community Water Supply. (1975) Health effects relat ing to direct and indirect reuse of waste water for human consumption. Report of an international working meet ing . Technical Paper 7. WHO-IRC, The Hague, The Netherlands .

2. Mil ler , D. G. and Short, C. S. Treatabi l i ty of River Trent Water. Symposium on Advanced Techniques in River Bas in Management: the Trent Model R e s e a r c h P r o g r a m m e . The Institution of Water Eng ineers , London (1973).

3 . Water R e s o u r c e s Board. (1973) Water R e s o u r c e s in England and Wales . HMSO, London, England.

4 . Department of the Environment. (1971) F i r s t Annual Report of the Steering Committee on Water Quality. HMSO, London, England.

5 . COST Project 64b Management Commit tee . (1976) A comprehens ive l i s t of polluting substances which have been identified in var ious fresh waters , effluent d i scharges , aquatic an imals and plants and bottom sed iments . Second edition EUROCOP-COST Secretar iat . The C o m m i s s i o n of the European Communit ies , B r u s s e l s , Belg ium.

6. Br idges , B. A. Short t e r m screening t e s t s for carc inogens . Nature 261, 195 (1976).

7. McCann, J. , Choi, E . , Yamasaki , E . and A m e s , B. N. P r o c . Nat. Acad. Sei . , US 72, 5135 (1975).

8. McCann, J. a n d A m e s , B. N. P r o c . Nat. Acad. S e i . , US 73, 9 5 0 ( 1 9 7 6 ) .

9. Green, M. H. L. , Muriel , W. J. and Br idges , B. A. Use of a s impli f ied fluctuation tes t to detect low l e v e l s of mutagens . Mutation R e s e a r c h 38, 33 (1976).

Health Aspects of Water Recycling Practices Prof. HILLEL I. SHUVAL

Director, Environmental Health Laboratory, Hebrew University-Hadassah Medical School, Jerusalem, Israel

ABSTRACT

Planned wastewater recycling for possible domestic consumption is under active study in a number of countries based on treatment by advanced biological and physico-chemical treatment processes. The question of whether conventional drinking water standards apply to the case of direct wastewater reuse is reviewed. The need for further intensive toxicological and epidemiological studies prior to such reuse is reviewed as well as special fail safe monitoring programs for finished water. It is not clear whether current day technology can meet all of the requirements for producing safe and wholesome drinking water from urban and industrial wastewater streams without considerable further study and evaluation.

INTRODUCTION

As water shortages grow in both arid and temperate areas of the world and as waste-water treatment required for the protection of the environment become more rigorous, the possibility of direct re-cycling of the highly purified wastewater effluents becomes economically attractive. Plans for direct re-use of wastewater processed by sophisticated, advanced treatment methods for domestic consumption are under active consideration in a number of countries including the United States of America, the Republic of South Africa, and the United Kingdom (Shuval, 1977).

The possible long-term health effects of consuming such water however is a major constraint in the development of such projects.

It has been pointed out that wastewater re-use for domestic consumption is de-facto being widely practiced indirectly in hundreds of cities presently drawing their drinking water supplies from heavily contaminated rivers. The lower reaches of many of the major rivers of the world such as the Rhine, the Thames, and the Mississippi carry heavy loads of treated and partially treated wastewater from urban and industrial sources. At times more than 50% of the flow of some rivers is made up of wastewater effluents. Dr. Packham (1973) has calculated the proportion of sewage effluent to total river flow at water abstraction points in the United Kingdom and shows that in many cases the effluent flow is greater than 20% reaching as high as 47% for the River Lee at Chingford. During the severe draught in the U.K. in the summer of 1976 the amount of wastewater flow in some major rivers approached 100% during certain periods. Cleary (1963) has estimated that the Ohio river carries about 18% effluent.

395

396 H. I. Shuval

Studies carried out in Europe and the United States have identified some 1000 organic and inorganic chemicals in wastewater effluent, raw river water and in drinking water supplies. Many of these chemicals are known for their toxic properties and some are known carcinogens.

In the light of these facts it is becoming apparent that the study of the health effects of wastewater reuse for domestic consumption is not limited to future planned projects for direct reuse but is even more vital for the hundreds of existing cases of indirect or covert reuse. It has been estimated that some 200 million persons throughout the world are presently consuming a significant percentage of residuals from wastewater sources.

The health effects of such practices have been studied only to a very limited extent so far.

A study in New Orleans (Page et al., 1975) compared rates of cancer among those drinking Mississippi River water with an adjacent control group consuming high quality ground water. The report indicates that a statistically significant relationship exists between cancer mortality rates in Louisiana and drinking water obtained from the Mississippi River. This is true for total cancer, cancer of the urinary organs, and cancer of the gastrointestinal tract. The authors conclude by stating that "while statistical studies cannot by themselves establish causality, this regression study supports the hypothesis that there is a link between carcinogens in drinking water and cancer mortality". The New Orleans study has been criticized on methodoligical grounds.

A group of experts who reviewed the findings felt that in the study drinking contaminated Mississippi water has not been conclusively implicated as the cause of the high rates of stomach and urinary tract cancer found in New Orleans. On the other hand, they felt that the subject is an important one to the health of millions of persons consuming such contaminated water and warrants careful study before any final conclusions are made.

The findings by Rook of the Netherlands (1974) and Jolly (1973) and Bellar (1974) of the U.S.A. concerning the formation of carcinogenic organohalides as a result of chlorination of water containing organic matter has added further concern as to the long-term health effects of consuming contaminated river water or treated wastewater. In many areas both sewage effluent and drinking water supplies are heavily chlorinated, prividing ample opportunity for such carcinogens to be formed.

The EPA (1975) carried a preliminary study to check the findings of the New Orleans study and to evaluate the health risk of consuming low levels of chloroform. Data for 50 cities have been compared with cancer mortality occurring in populations served by these water utilities. The study, although preliminary, has indicated a statistically significant correlation between the cancer mortality for all anatomical sites and both sexes combined in the years 1969-71 with the chloroform concentration in samples collected in the Spring of 1975. Such a correlation was not noted with total mortality or with the sum concentration of the four trihalomethanes in drinking water.

The question as to the degree of risk to human health associated with relatively low levels of toxic chemicals in drinking water is a moot one. It is today accepted by the scientific community that the majority of human cancers are environmental in origin. Mutagenic and teratogenic effects from chemicals in the environment are well known.

Some feel that exposure to such chemicals which often appear at higher doses

Health Aspects of Water Recycling Practices 397

in food may be a more serious environmental hazard than water. However, on the other hand no single food product is ever as universally consumed as drinking water, to which the total population at all ages and in all stages of health and disease are exposed daily throughout their entire life span.

Although there is still no universal agreement as to the health risks associated with the exposure to very low levels of carcinogens over long periods of time, there is a growing body of scientific opinion who support the contention that, "No level of exposure to a chemical carcinogen should be considered toxicological insignificant for man." (Epstein, 1974).

The universality of water as a potential long-term source of exposure to even low levels of environmental contaminants, many of them now clearly identified as being carcinogenic or mutagenic, can no longer be neglected. Particularly since some 200 million people throughout the world daily consume water from surface sources contaminated with ever growing amounts of urban and industrial waste water.

The growing concern as to possible health risks associated with consuming surface water contaminated with a myriad of known and unknown organics applies a fortiori to the question of the possible health risks of direct recycling of waste water for domestic consumption.

WHAT CHEMICAL STANDARD APPLY TO WASTEWATER REUSE FOR DOMESTIC PURPOSES?

At this point, it is appropriate to ask whether our knowledge of the toxicological and epidemiological implications of wastewater renovation for domestic consumption is sufficient. Is the public health adequately protected if processed municipal and industrial wastewater can be brought in line with todays conventional drinking water standards? Were these standards conceived with such a possible application in mind? If not, what remains to be done prior to giving the final green light to total direct wastewater reuse for all purposes including human consumption? We shall attempt to discuss these questions as well as propose possible approaches to answering some of the yet unanswered questions.

The U.S. Public Health Service (1962) Drinking Water Standrads list 20 chemical parameters, only 9 of which serve as absolute grounds for rejecting a supply as unsafe for human consumption. The World Health Organization (1971) Drinking Water Standards contain a few more chemical parameters which may serve as grounds for rejecting a supply. EPA proposed drinking water standards will expand the list a bit further. None of these widely known and accepted standards list more than a few synthetic organic compounds, despite the fact that hundreds of such chemicals may find their way into municipal and industrial wastewater and many of them are known for their potential deleterious effects on human health. For that matter, neither do these standards exhaust the list of potential inorganic toxicants that may be found in industrial wastes.

It must be recognized that conventional drinking water standards were originally based on the assumption that water for human consumption would generally be drawn from groundwater sources or from the "best available" protected uncontaminated surface water sources and that the limited number of chemical parameters included were adequate for most situations. This assumption is rarely true for most surface supplies of today. Can standards be developed to cover the wide range of contaminants that are actually found in wastewater destined for domestic consumption?

Several hundred threshold-limit values for such industrial chemicals in air have

398 H. I. Shuval

been established under food regulations. The drinking water standards of the USSR Ministry of Health (1961) are the first to have recognized the scope of the problem and now include some 400 chemical parameters with more under study.

The mere comparison of the quality of the final effluent against those standards currently listed in the USPHS, WHO, or EPA drinking water standards leave too many questions unanswered to be accepted as adequate evidence that such an effluent is completely safe from the public health point of view.

The WHO International Working Group (World Health Organisation, 1975) concluded "...that conventional drinking water standards alone cannot provide a sufficient basis for the health evaluation of reused water for domestic consumption."

Not enough is known about the true identification of the residual microchemical pollutants both inorganic and organic which of course, vary widely from one situation to the next, depending on the nature of the industrial wastes that enter the sewerage system.

TOXICOLOGICAL EVALUATION OF RENOVATED WASTEWATER

There are two possible approaches to the toxicological evaluation of renovated wastewater to be used as drinking water (Shuval and Gruener, 1973). The first would require the establishment of maximum allowable concentrations or limits for each of the potentially hazardous chemicals that may be found in renovated wastewater. The approach that has been developed by toxicologists in setting tolerance limits for food additives and chemical contaminants in food has been to establish acceptable daily intake (ADI) levels for man. Although much of the basic data may be available to assist in setting such standards for water, important information is still missing, (World Health Organization, 1973).

The toxicological evaluation of chemicals found in the environment cannot be simplified to take into account acute or subacute effects alone. Today, such evaluation must include effects from long-term exposure and studies for carcinogenicity, mutagenicity, teratogenicity, and various biochemical and physiological effects. Even if the specific toxicity of defined industrial and agricultural chemicals is established, the possible toxic effects of their breakdown products may be more difficult to determine. Natural biodegradation or specific treatment processes such as chlorination may lead to the development of new compounds having toxic properties quite different from those of the parent compound. Work to identify these breakdown products and to study their toxic effects is required.

Another factor complicating the toxicological evaluation of heavily polluted water or renovated wastewater is the combined and possible synergistic effect resulting from the exposure to a mixture of toxic and nontoxic chemicals. Increased toxic impact to such combinations is known to occur under certain circumstances and the case of renovated wastewater must take into account such possibilities.

Although much is to be gained by establishing proper tolerance levels for many of the known toxicants that might appear in renovated water, this approach will take a long time to develop and even then will not cover all possible toxic effects as pointed out above.

For these reasons, it is felt that a second approach is required. A full toxicological evaluation should be carried out on the actual finished renovated water intended for human consumption with its real mixture of known and unknown residual chemicals remaining after treatment.


Long-term feeding experiments with more than one species of experimental animal should be required, as well as other more rapid toxicological screening tests using bacteria or cell cultures. Such studies would include testing concentrates of the residual chemicals in the final processed water, as well as the normal unconcentrated effluent. Concentration techniques used must avoid being selective as is the case with activated carbon and must not lead to the breakdown of the chemicals involved by overly harsh treatment, such as high temperature distillation.

Most governments require the full toxicological evaluation of any new drug or food additive before allowing its commercial use. The requirements for evaluating renovated wastewater with its many and often complex unknowns should be, at least, as rigorous. WHO reports on this matter (World Health Organization, 1973, 1975) have emphasized the need for complete toxicological evaluation of the actual water planned for reuse as an essential step in evaluating the safety of such reuse.

BUILDUP OF DISSOLVED SOLIDS IN RECYCLING

In multiple recycling of wastewater, there will be a buildup of those dissolved solids not removed or only partially removed by the wastewater treatment plants unless specific demineralization processes are included. However, the buildup in concentration will not be infinite, since there will be fresh make-up water added in each cycle to compensate for water losses that do not appear as sewage flow. In most cities, these losses normally range from 10-20%.

Thus, if we assume that in each recycle 90% of the water input into the community appears as wastewater which will be processed for recycling, the concentration say, of, sodium chloride will increase with recycling until it reaches an equilibrium 10 times greater than its original concentration in the wastewater. The same might be true for certain refractory organics of potential public health danger. Under such conditions, partial demineralization would be necessary to keep dissolved inorganics at acceptable levels. The problem of toxic trace organics and inorganics may be more difficult to deal with since these substances may not all be removed with equal effectiveness by some demineralization processes which may be selective. The buildup of such toxic materials on multiple recycling would be very undesirable. Until such time as complete information is available on the removability of the various toxic organic and inorganic trace elements that may appear in the wastewater stream, it might be prudent to reduce the possibilities of buildup by providing additional dilution from freshwater sources. For example, with a reuse factor of only 30%, the maximum concentration at equilibrium of a chemical not removed at all will be only 40% greater than the original concentration in the wastewater stream, while with a reuse factor of 50%, equilibrium will be reached with the concentration of the refractory compound of twice its original concentration. Such dilution with freshwater would normally avoid the need to include an expensive demineralization step in the treatment and would provide an additional safety factor against the buildup of compounds whose removability by demineralization or other processes may not as yet be known.

EPIDEMIOLOGICAL EVALUATION

No matter how thorough a toxicological evaluation is made, there always remains the problem of extrapolating the findings with laboratory animals to fit the human situation. With drugs of potentially great medical importance, human trials are held after completion of the toxicological evaluation. In the case of new food additives that are usually less essential to human welfare than drugs, negative findings in the toxicological evaluation do not automatically mean the new chemical will be allowed for use in human food. It must be demonstrated that the chemical will make a significant contribution to improving the quality or

400 H. I. Shuval

preservation of the foods in which it will be used. In water-short areas, water renovation can often be justified as being of potentially great importance to human welfare as well.

A considerable body of information has been built up concerning the effect on human health of many of the environmental contaminants discussed here as a result of direct exposure of humans under various industrial situations. Further information has been gained by the accidental exposure of humans to certain toxic materials.

If wastewater renovation for domestic consumption is ever to become widely accepted, there will be a need at some stage to carry out a full-scale epidemiological evaluation of the impact of such reuse on the health of the population exposed. It may be difficult to choose an appropriate population group for such a study, but to the extent that certain communities in water-short areas have already gone ahead with wastewater reuse for drinking water, every effort should be made to carry out a thorough epidemiological evaluation. Such a study should include baseline health evaluation of a sample population before the introduction of renovated water and then a follow-up of the same group, as a panel study, over a 5- or 10-year period. Such opportunities will be few and far between, and every effort should be made to gain as much data from each case as is possible.

A promising alternative to such a study with a population exposed to planned wastewater reuse would be a series of studies of populations exposed to indirect or unintentional wastewater reuse. Such population groups are easier to identify than might be imagined, since many millions of people throughout the world are, in fact, consuming renovated wastewater everyday and have been doing so for years. Millions of people consume water from rivers which, at times of low flow, may contain as much as 40-50% of industrial and municipal wastewater.

It must be recognized that unintentional and, in many respects, uncontrolled reuse of wastewater is now very widely practiced and provides a basis for evaluating the health impact of such use as well as the expected impact of fully engineered and carefully controlled direct wastewater reuse of the type under discussion.

Although such prospective epidemiological studies are expensive and take many years to complete, it is essential that they be made even if planned direct wastewater reuse were not under consideration. Such studies are essential to evaluate present environmental exposure from consuming water from polluted sources, which will become even more polluted in years to come. Epidemiological studies may shed light on the need to make major improvements in present-day water treatment technology, which has been demonstrated as being relatively ineffective in removing many of the refractory organic toxicants that appear in increasing quantities in polluted water. The findings will also be of vital importance in planning future wastewater renovation programs, where it can surely be expected that treatment trains of demonstrated efficiency will be utilized to remove potentially hazardous chemicals or pathogenic microorganism to the lowest possible levels.

MONITORING THE REUSE OF WASTEWATER FOR DOMESTIC CONSUMPTION

The nature of a monitoring or quality-control program for products produced for human consumption should vary according to the degree of risk to health involved. Conventional water supply monitoring programs have in the past assumed that the product is basically a safe one and that it can be supplied to the consumer directly after processing without waiting for the results of quality-control tests. Bacteriological test results are usually available 24-48 hours after


sampling, while routine tests for toxic chemicals, when they are made, may be available only after many days. The water tested has usually been consumed by the population by that time.

Drugs and food additives and many processed foods are tested routinely and are released for use in batches only after the test results indicate no positive findings.

In the case of a plant for renovating wastewater for human consumption, it would appear necessary to require a more rigorous monitoring and quality control regime than that currently practiced by the water supply industry. Technical breakdowns and human failures at such a plant might lead to major health hazards. It would not be illogical to require that renovated water be fully tested and certified as safe before its release to the general water-supply system. With improved bacteriological techniques, results can be obtained in under 24 hours as can the results of most of the important chemical tests, many of which can be automated. Ways of carrying out rapid toxicological evaluation of the finished water with bioassay techniques should be developed. For the moment, virus assays require at least 5 days for completion, but, here too, more rapid assay techniques are under study and may become available.

Renovated water could be produced and held in batches until completion of the quality-control tests, before being released. This will add additional costs to wastewater renovation plants, but the additional safety obtained would justify the expenditure. Certainly, such precautions should be practiced in all early plants until it can be demonstrated that less stringent quality-control measures are adequate.

Many might agree that the proposed monitoring regime should be applied to any case where heavily polluted surface water is the source of drinking water supplies. Such supplies may be an even greater risk than planned direct reuse programs.

POLICY CONSIDERATIONS

The approach presented here concerning water reuse for domestic consumption may appear to be overly cautious and place too heavy a burden on future wastewater renovation programs. In answer, it must be stated that criticism of current drinking water standards applies as much to any case where polluted surface water serves as a source of drinking water as to the special case of direct wastewater renovation. In fact, indirect, unplanned wastewater reuse may well be a greater risk than planned direct reuse, which would include treatment processes more capable of coping with the organic pollutants found in wastewater. Unplanned or covert wastewater reuse is far too widely practiced today, with too few controls, to allow one to feel complacent.

However, planned direct wastewater reuse for domestic consumption carries a heavy responsibility with it, since it involves full engineering and health responsibility from the beginning to the end, without the intervening hand of "nature." The fact that nature provides little protection in heavily polluted rivers whose self-purification capacity is overtaxed gives little justification for a similar lax approach in a planned direct reuse project. In such a project, the designers, operators, and health authorities who must give their approval must carry the full responsibility of any adverse health effects which may result, even if it can be shown that communities consuming polluted surface water are exposed to equal or greater risks.

Many such communities are indeed exposed to undesirable health risks and, thus a subsequent equal tightening up of standards, treatment procedures, and quality

402 H. I. Shuval

control for all cases of wastewater reuse whether direct or indirect is fully justified.

It still remains to be demonstrated that water treatment technology can overcome the many problems involved in processing wastewater with its many complex components and rapid fluctuations in quality to achieve a uniform end product meeting the health requirements for wholesome and safe drinking water. However, further work in this direction will be very important, whether it be applied to direct water reuse projects or to the more urgent and widespread cases of indirect, covert wastewater reuse that exists so widely today in communities drawing water from the polluted lower reaches of the great rivers of the world.

Another consideration that cannot be overlooked is that of public attitudes toward water reuse for domestic purposes. The strong public opposition which, in many cases, thwarted efforts to introduce fluoridation despite strong technical evidence and the support of the scientific community is an illustration of an aroused public. Water reuse for domestic consumption may not be easily accepted by the public, even if all the precautions outlined above are taken. The study of Bruvold and Ward (1972) in 10 towns in California indicated that out of 25 forms of possible wastewater reuse only 11 would be likely to receive no public opposition. These include such items as golf course irrigation, commercial air conditioning, and hay, alfalfa, or orchard irrigation. Over 50% of those interviewed opposed reuse for domestic purposes. Any planned programs for reuse must give careful consideration to this question from the very beginning, or they- may find years of scientific and technical effort vetoed by the public.

In the final analysis, direct planned wastewater reuse for human consumption may well become feasible through the development of advanced wastewater treatment systems with a demonstrated fail-safe capability of removing the hundreds of potentially toxic inorganic and organic chemicals that appear in today's wastewater streams. A major combined effort of developing appropriate advanced technology and health effects evaluation will certainly be required to achieve this goal.

REFERENCES

Bellar, T.A., Lichenberg, J.J. and Kroner, R.C., The occurrence of organohalides in chlorinated drinking waters, J. Amer. Water Works Assoc. 66, 703-706 (1974).

Bruvold, W.H. and Ward, P.C., Using reclaimed wastewater-public opinion, J. Water Pollut. Contr. Fed. 44, 1690-1696 (1972).

Cleary, E.J., Horton, R.K. and Boes, R.J., Re-use of Ohio River water, J. Amer. Water Works Assoc. 55, 683-686 (1963).

E.P.A. (1974) Preliminary assessment of suspected carcinogens in drinking water: Report to Congress. U.S.E.P.A. Washington, D.C., December, pg. 30.

Epstein, S.S., Environmental determinants of human cancer, Cancer Res. 34, 2425-2435 (1974).

Jolly, R.L., Chlorination effects on organic constituents in effluents from domestic sanitary sewage treatment plants, Publ. no. 55, Environ. Sei. Div., Oak Ridge Nat. Lab., Oak Ridge, Tennessee (1973).

Page, T., Harris, R. and Epstein, S.S., Drinking water and cancer mortality in Louisiana, Science, 193, 55-57 (1975).

Packham, R.F., Potable water from sewage effluent. Symposium Sewage Effluent as a Water Resource (14-15 Nov. 1973) Inst. of Public Health, Eng., London.

Rook, N.J., Formation of haloforms during chlorination of natural waters, J. Soc. Water Treat. Exam. 23, Part 2, 234-243 (1974).

Shuval, H.I., The public health significance of trace chemicals in wastewater utilization, Bull. W.H.O. 27, 791-799 (1962).


Shuval, H.I., Water Renovation and Reuse, Academic Press, New York, 463 pgs. (1977).

Shuval, H.I., and Gruener, N., Health considerations in renovating wastewater for domestic use. Environ. Sei. Technol. 7, 600-604 (1973).

U.S. Public Health Service, "Drinking Water Standards", U.S. Dept. of Health, Education and Welfare, Washington D.C. (1962).

U.S.S.R. Ministry of Health, "Standards for Drinking Water", Moscow (1961). World Health Organization, "International Standards for Drinking Water", World

Health Organ., Geneva (1971). World Health Organization, Reuse of effluents: Methods of wastewater treatment

and health safeguards, Tech. Rep. Ser. 517, World Health Organ., Geneva (1973). World Health Organization, "Report of the International Working Meeting on Health

Effects relating to Direct and Indirect Reuse of Wastewater for Human Consumption", International Reference Center for Community Water Supply, Amsterdam (1975).

In Vitro Mutagenesis and Carcinogenesis Testing of Residual Organics in Drinking Water

J. C. LOPER*, D. R. LANG*, C. C. SMITH*, R. S. SCHOENY*, F. C. KOPFLER** and R. G. TARDIFF**

^University of Cincinnati Medical Center ** National Environmental Research Center, U.S. EPA, Cincinnati, Ohio

ABSTRACT

Reverse osmosis was employed to obtain residual organics in sequential samples of drinking water from 5 cities representative of different sources. Samples from all the cities produced dose-related mutagenesis using the Ames Salmonella strains. Microsomal activation did not enhance these effects. Distribution of mutagenic activity per mg into hexane, ethyl ether, and acetone soluble subfractions was es-tablished on sequential samples of 2 cities. Mutagenic patterns were generally constant for samples taken at 3 month intervals, and showed city-specific character-istics. In vitro morphological transformation of clone 1-13 of Balb/3T3 was in-duced by the one sample tested.

INTRODUCTION

Assessment of health effects of organic compounds in drinking water is a formidable problem. The known constituents are relatively few, numbering a few hundred chemicals among mixtures postulated to contain 3000 or more yet to be identified compounds. Most are present at very low levels. Toxicological analysis of such mixtures requires some type of initial concentration procedure.

This study was begun to test the applicability of reverse osmosis in the concentra-tion of drinking water organics. Sequential samples have been prepared from drinking water of cities representative of United States municipal water sources. This paper reports studies of such residues using two in vitro assays, the Salmonella/ microsome system, and mammalian cell transformation.

MATERIALS AND METHODS

Test Samples.

Residual organics were generated for USEPA by Gulf South Research Institute using the reverse osmosis procedure described by Kopfler, et_ aj_. (1). The overall pro-cedure is summarized in Fig. 1. Twenty percent aliquots were taken at each step for subsequent GC/MS analysis. Each water sample was processed sequentially to produce 1) the reverse osmosis concentrate organic extract (R0C-0E) and 2) an ethanol-soluble eluate obtained from a macroreticular resin XAD-2 column (XAD eluate).

405

406 J . C. Loper, et dl.

WATER SAMPLE

R.Q Cellulose Acetate

R.O. Nylon

Cellulose Acetate Concentrate

Nylon Concentrate

i Pentane pH7

\ Methylene

Chloride pH7

i r Methylene

Chloride pH2

1 r Υ Λ Π - Ο

1 /\ h\\

pH \z

80%

20% sample

\

Ethanol Elution

ROC-OE

80% 1

2 0 % sample

-— * /

/

Y Λ π

Eluate

Ethanol -— Elution

Pentane pH7

f

Methylene Chloride pH7

\ Methylene

Chloride pH2

\ r ΥΛΠ-9 I

PH 2 J

T discard discard

Fig. 1. Origin of reverse osmosis concentrate-organic extract (ROC-OE) and XAD eluate samples.

In Vitro Mutagenesis and Carcinogenesis Testing 407

TABLE 1 Cities selected for Extraction and Bio-Assay of Residual Organics in Drinking Water

Type of Type of City Water Supply Raw Water

^liami, FL |New Orleans, LA bttumwa, IA Philadelphia, PA Seattle, WA

Ground Surface Surface Surface Surface

Uncontaminated* | Industrial Wastes ] Agricultural Runoff 1 Municipal Wastes | Uncontaminated* |

*Refers to no known contamination from municipal, agricultural, and industrial wastes -- but contamination presumably from decomposition products of natural origin.

These procedures were applied repetitively to drinking water samples from 5 U.S. cities (Table 1). Six samples were prepared from each city over a 15 month period. The sample sets were designated for use in a protocol of In vitro and jji vivo testing procedures described elsewhere (2); city samples 1 and 2 were available for short term tests of the type presented in this paper. It was proposed at the outset that the ROC-OE residues should be obtained in approximately one gram amounts. Volumes of water processed and yields for samples most pertinent to this paper appear in Table 2. Following preparation of the samples identified as la, a Donnan softening loop was included in the concentration process, reducing the content of divalent cations in the concentrate by exchange with sodium ions (1).

TABLE 2

pap water processed, liters

Date of collection

Total organic carbon in water (ppm)

Reverse osmosis concentrate, or-ganic extract |(ROC-OE), g

XAD eluate, g

Total yield, g

Vield of Organics from Drinking Water

City and Sample New Orleans

la lb 2

7994 7812 6624

10/75 10/75 1/76

2.0 2.0 1.7

1.18 0.99 1.45

4.61 5.56 6.41

5.79 6.55 7.86

Miami

la

1999

11/75

6.4

1.08

7.18

8.26^

2

2271

2/76

6.4

0.96

8.78

9.74J

408 J. C- Loper, et at.

A portion of each ROC-OE and XAD fraction was further fracionated by sequential extraction with hexane, ethyl ether, and acetone according to a method of R.G. Mel-ton (U.S. EPA report, Cincinnati, Ohio 1976). These ROC-OE and XAD fractions, and their respective subfractions were used in the bioassays reported here.

Bacterial Mutagenesis Testing

The Salmonella/microsome system has been described by Ames, et al_. (3), who pro-vided strains TA1535, TA1538, TA98 and TAIOO. Strain TA1535 is an indicator of missense mutagenesis, while TA1538, and TA1538 with the plasmid pKMlOl, renamed TA98, are detectors of frameshift mutagens. Similarly, TAIOO is TA1535 (pKMlOl). The presence of pKMlOl increases the sensitivity of these strains to mutagenic com-pounds and with TAIOO, in particular, may alter specificity to include detection of frameshift as well as missense mutagens. Promutagen activation on the petri plate used S-9 mixtures prepared from livers of rats induced with a PCB mixture, Aroclor 1254. Activation of 2-aminoanthracene served as the positive control. Based upon initial results, an assay protocol was applied as follows. Samples were dissolved in dimethyl sulfoxide (DMSO) and delivered in volumes of 0.01 to 0.30 ml/ plate. All tests involved duplicate platings of 5 doses over a 30-fold dose range for mutagenesis of TA98 and TAIOO in the absence of S-9 mix. This assay was then repeated, with dose adjustments as appropriate, in an expanded protocol which in-cluded the addition of the microsome system, and optionally, the strains TA1535 and TA1538. Generally a previously characterized direct acting mutagenic test sample was included as an additional control. This procedure provides data on the muta-genic potential of a sample while retaining much of it for further fractionation experiments. Dose responses were recorded in specific activity units of net rewer-tant colonies/mg of sample as determined from linear regression plots generated with a computer-plotter.

Cell Culture

Clone 1-13 of Balb/3T3 was obtained from Dr. Takeo Kakunaga (N.I.H., Bethesda, MD). Cells were grown in 75 cm^ tissue culture flasks (Lux Scientific Corp., Biolabs Inc., Northbrook, IL 60062) using antibiotic-free Eagle's minimum essential media (MEM) supplemented with 10% heat inactivated fetal calf serum (both from Grand Island Biologicals Co., Grand Island, NY). Cells were then incubated at 37C in a humidified incubator with an atmosphere of 5% CO2 in air. Cells were passaged once per week using 0.25% trypsin and maintained at sub-confluence.

Transformation Assay

The experimental conditions used are essentially those described by Kakunaga (4). A subconfluent culture of cells was selected and then plated at 1 x 104 cells per 60 mm tissue culture dish in 5 ml media and incubated for 18-24 hours. For each experimental point 10 dishes were prepared as well as 10 dishes for controls. Appropriate concentrations of carcinogen or water sample in 10 yl DMSO were then added to the medium already present on the cells; control dishes received 10 yl DMSO alone. All samples were added to cultures under red light to prevent photo-oxidation. Cultures were incubated for 72 hours; the medium then was removed and replaced with fresh medium. The cultures were refed twice weekly. Usually, foci could be identified readily by the fourth week. For counting, the plates were rinsed with phosphate buffered saline, pH 7.2 (PBS), fixed with methanol, and stained with Giemsa; areas of piled up cells growing in a disorganized, criss-cross fashion were quantitated as foci. Clones selected from transformed foci as well as cells from "normal" areas of treated plates were tested for their plating effi-ciency in soft agar by the method of MacPherson and Montagnier (5).

Cytotoxic Effects


Toxicity of each carcinogen or water sample was also determined using Balb/3T3 cells. Cells were plated at a concentration of 200 cells per 5 ml MEM per 60 mm dish. Test samples were added as for the transformation assay in 10 μΐ volumes, with controls receiving 10 yl DMS0. Exposure was for 72 hours after which time the medium was removed and fresh MEM was added. These cultures were fed twice weekly for 10-14 days, and then fixed and stained with Giemsa. Colonies on each plate were counted and the plating efficiency was determined by dividing the average number of colonies per test group by the total number of cells plated.

RESULTS

To date, the Salmonella mutagenesis procedure has been applied to ROC organic ex-tracts and XAD eluates, and their hexane, ethyl ether, and acetone subfractions from at least one sample of each city. Mutagenic activity was detected for R0C-0E from each city, the TA98 response ranging from none to 10-fold the average spontan-eous background, and the TA100 response generally equalling a 2-fold response. However, more useful comparisons are possible in those cases where repeat samples from the same city have been analyzed. New Orleans la and lb samples were collected at the same time, with sample 2 being generated 3 months later. Similarly, Miami la and 2 were collected with 3 months intervening. Both la samples were concen-trated without the sodium ion-divalent cation exchange loop used subsequently. Tables 3 and 4 list the mutagenic activity for TA98 and TA100 per mg of these 5 parent samples and their subfractions, and also the weight distribution of organics among the subfractions. Many of the samples showing mutagenicity gave linear activity plots. Mutagenic activity for TA1538 was usually similar to that for TA98 while TA1535 was unaffected in several cases where TA100 showed a response. An example of this pattern using Miami 2 R0C-0E is shown in Fig. 2. Several fractions appeared toxic and gave non-linear, variably mutagenic responses (see Tables 3 and 4).

40

_ J

CO

o

I— cxl LU

>

20 \

400

200

TAIOO

750 1500 0

yG/PLATE

750

TAI535

1500

Fig. 2. Strain speci f ic mutagenesis of Miami 2 ROC-OE (see Table 2)

TABLE

3 Fractionation

of N

ew O

rleans S

amples - Distribution of W

eigh

t and

of

Mutagenic

Activity for

TA98

and T

A100

Starting

Material

and

Sub-

fractions

ROC-OE

1 hexane (CH)

ethyl

ether

(CE)

acetone

(CA)

XAD

1 hexane (XH)

1 ethyl

ether

(XE)

[ acetone

(XA)

♦Percent f

igures f

or

insoluble

residue.

**Average from 2 expe

no d

ose

response;

n

mg a

nd P

ercent o

f To

tal*

la

lb

2

603(

51)

262(

27)

649(

45)

3(1)

22

(6)

67(1

0)

437(

83)

304(

79)

524(

79)

64(1

2)

57(1

5)

71(1

1)

1227(27)

1749

(32)

18

17(2

8)

23(2

) 42

(2)

38(2

)

548(

42)

435(

24)

785(

43)

611(

47)

1225

(66)

76

2(42

)

TA98

165±12

no d

.r.

65±2

8

158±

4*

Net

la

TA10

0

250±3

no d

.r.

144 m

Revertant lb

TA98

187±

12*

no d

.r.

52*

162*

Colo

nies

/mg*

*

2 TA

100

TA98

283±

30*

76±33

no d

.r.

m

107*

57±0

no d

.r.

183±

14*

TA10

0

225±

59 1

m j

172±

3 J

V

n.d.

no d

.r.

m

no d

.r.

n.d.

no d

.r.

m

no d

.r.

no d

.r.

m m m

no d.r.

m

v m*

m m*

*

hexa

ne,

ethyl

ethe

r, a

nd a

ceto

ne s

ubfr

acti

ons

are

relative t

o total

weig

ht r

ecovered inc

riments of values taken from linear portion of

dose response

plo

ts.

Othe

r symbols: no d.

r.,

d.,

not

done

; *

appa

rent

toxic

effe

cts;

m,

marginal mu

tage

nic

resp

onse

; v, widely

variabl

m 1

m*

m*

m*

luding

e

r1

o ft

dose r

esponse.


TABLE 4 Fractionation of Miami Samples - Distribution of Weight and of Mutagenic Activity for TA98 and TA100

Starting Material and Sub-fraction

ROC-OE

CH

CE

CA

XAD eluate

XH

1 XE

1 XA

mg and Percent of Total* la 2

677(63)

95(11)

536(65)

191(23)

438(46)

22(5)

275(67)

109(27)

2765(39)

131(5)

1497(51)

1000(34)

3081(35)

22(1)

1454(63)

800(35)

Net Revertant Colonies/mg**

TA98

42±5

no d.r.

52±7

64

la TA100

137±11

347±75

166±

no d.r.

TA98

17±4

no d.r.

20±3

38±2

2 TA100

179±2

800±28

288±13

142±3

no d.r.

137±22*

19±6

no d.r.

no d.r.

821±136*

112±20

no d.r.

no d.r.

no d.r.

no d.r.

no d.r.

no d.r.

1226±5

49±2

no d.r.

♦Percent figures for hexane, ethyl ether, and acetone subfractions are relative to total weight recovered including insoluble residue ♦♦Average from 2 experiments of values taken from linear portion of dose response plots. Other symbols: no d.r. = no dose response; ^apparent toxic effects.

Certain samples showing little or no mutagenesis but no toxicity were tested for their effect on assays of known compounds. Such an experiment testing the effects of New Orleans la XAD eluate, or of its ethyl ether subfraction (XE), upon muta-genesis by 4-nitroquinoline-N-oxide, or by methyl methanesulfonate is shown in Fig. 3. To date no qualitative supression or amplification of mutagen detection has been observed.

Systematic addition of the microsome activation system produced little or no en-hancement of mutagenic effect in any of the samples tested. A representative res-ponse is shown in Fig. 4. A study was made of the effect of certain concentrates upon the activation of a known promutagen. Addition of Ottumwa 1 ROC-OE to benzo-[a]pyrene and the direct action system gave what appeared to be additive effects of benzo[a]pyrene and the direct acting mutagens in the residue fraction (Fig. 5). Analogous effects were obtained with 2-aminoanthracene.

Table 5 shows the results of a transformation experiment using the known carcinogen 3-methylcholanthrene (MCA) as a positive control and New Orleans lb ROC-OE. The water sample showed transformation rates higher than the control at concentrations ranging from 10 to 100 yg/ml. When these data are expressed in terms of foci/sur-vivor, which takes into account the toxicity of the sample, a rate of 7.4 x 10"3

was obtained with 100 yg/ml of ROC-OE. This value approached that of 1.1 x 10-2 foci/ survivor observed with 2.0 yg/ml of MCA. When MCA and ROC-OE are added

412 J, C, Loper, et dl.

together it is evident that the presence of such a complex mixture of organics did not preclude the demonstration of the MCA transformation as the rate is not significantly different than that of MCA alone.

300 r

CO

o o eo

300

NQNO

NQNO +XE

BO

MMS

MMS + XAD

0 I 2 yG OF MUTAGEN/PLATE

1500

Fig. 3. Effect of New Orleans la XE, 1000 yg/plate, and XAD, 500 yg/plate (see Table 3) on mutagenesis by 4-nitroquinoline-N-oxide (NQNO) and methyl methanesulfonate (MMS).

o

to

90

25

TA98

y ^

1

O

S Δ

1

400 r TA100

+ S9 -S9

200

-S9

S9

500 1000 0 yG/PLATE

500 1000

Fig. 4. Mutagenesis of Miami la CE (see Table 4) in the absence and presence of the microsomal activation system (S9)


I — Cd

1000

500

/ °

A

1 ^_

1

▲

o \

1 1

V " '——--*?

X O

BAP

BAP + OTT ROC

10 yG OF BENZO[A]PYRENE (BAP) ON THE PLATE

Fig. 5. Effect of Ottumwa 1 ROC-OE, 750 yg/plate, upon benzo[a]pyrene activation

Fig. 6 gives an example of the appearance of both normal cells and the edge of a typical focus obtained with this sytem. These cells have been fixed and stained with Giemsa.

The focus shown was obtained by treatment of the cells with 10 yg/ml benzo[a]pyrene; its morphology is characteristic of the foci we include in our quantisation. Clones obtained from these foci show the typical characteristics of transformed cells including disoriented growth, decreased adhesiveness, increased saturation density, and increased plating efficiency in soft agar.

DISCUSSION

Results of this ongoing study clearly indicate that organic residues isolated from drinking water by the reverse osmosis procedure produce bacterial mutagenesis. These residues and their organic solvent subfractions show different patterns of mutagenic effects. For the cities so far studied, the patterns are relatively stable for samples taken at 3 month intervals. One sample was shown to transform Balb/3T3 cells producing cells which exhibit several properties of tumor cells jn_ vitro including enhanced plating efficiency in soft agar.

Analysis of complex mixtures such as these water concentrates presents obvious problems even in the use of the relatively simple in vitro tests applied here. Although the possible mixtures of mutagenic and toxic components add variability

TABL

E 5

T

ran

sfo

rmat

ion

o

f 3T

3-11

3 c

ell

s by

New

Orl

eans

RO

C-O

E

Sam

ple

Co

ntr

ol

MCA

2.0

ug/m

l

New

Orl

eans

0.2

ug/m

l 2.

0 "

10.0

"

50.0

"

100.

0 "

MCA

2.0

ug/m

l +

ROC-

OE

II

II

II II

II II

II II

R0C

-( 0,

2.

10,

50,

100,

3E

.2

.0

.0

.0

.0

/di

tota

l /

wi is

hes

ith

d

ish

es

/ fo

ci

12/2

9/9

8/2

8/0

8/6

7/

4 12

/5

6/6

6/6

9/9

9/9

6/5

tota

l #

foc

i

4

118 2 0 11

8 16

75

77

91

93

21

foc

i/

dis

h

0.33

13

.1

0.25

0.

00

1.38

1.

14

1.33

12.5

12

.8

10.1

10

.3

3.5

abso

lute

PE

16.6

11.8

14.0

16

.6

15.4

9.

5 1.

8

14.2

13

.5

14.1

7.

7 1.

7

rela

tiv

e PE

100 71

85

100 93

58

11

86

82

85

47

11

foc

i/

surv

ivo

r

2.0

1.1

1.8

9.0

1.2

7.4

8.8

9.

5 7.

2 1.

3 2

.1

x 10

"4

x 10

"2

x 10

"4

0 -4

x

10 Z

x

10"^

x

10" J

x 1

0'^

x

10":?

x

10

",

x 1

0",

x

10"^

10

ce

lls

wer

e p

late

d

and

allo

wed

to

at

tach

o

vern

igh

t.

The

foll

ow

ing

da

y sa

mpl

es w

ere

adde

d i

n

10 u

l DM

SO a

nd a

llow

ed

to

rem

ain

on

th

e c

ell

s fo

r 72

ho

urs

. S

ampl

es

wer

e re

mov

ed,

the

ce

lls

rin

sed

wit

h

PBS

and

re

fed

. C

ult

ure

s w

ere

refe

d

biw

eekl

y fo

r 30

day

s,

fix

ed

, st

ain

ed w

ith

G

iem

sa,

and

qu

an

tita

ted

.

o


Normal 3T3 1-13 cells 3T3 1-13 cells transformed with 10 ug/ml benzo[a]pyrene

Fig. 6. Examples of normal and transformed 3T3 1-13 cells.

416 J. C. Loper, et dl·

to the assays, and probably alter overall dose response curves, some conclusions appear to be straightforward.

1) The reverse osmosis procedure yields relatively large amounts of residual organics, sufficient both for short term assays and for further fractionation.

2) Several of the subfractions obtained by solvent extraction induce apparently simple mutagenic responses. Some revert TA1538 and TA98 but not TA1535 unless it contains the pKMlOl plasmid (strain TA100), indicating that these subfractions contain primarily frameshift mutagens. Combined use of specific bacterial testers in conjunction with further chemical purification should lead to eventual identifi-cation of the active compounds.

3) Indications of mutagenic or carcinogenic activity in these mixtures can be compared with similar data being generated with known constituents of drinking water (6). This should make possible the detection of comutagenic, cocarcinogenic, or antagonistic activities of components in the mixtures.

There is a continuing need for reliable methods in assessing the presence of non-volatile organics in drinking water and drinking water sources. The degree to which organic residues generated by this reverse osmosis procedure appear to be re-presentative of the dilute mixture in the sample water is discussed elsewhere (1).

Considerable attention is being paid to direct XAD adsorption as a concentration procedure (7,8). Residues in water processed in parallel by reverse osmosis and by direct adsorption using XAD resins should be compared for the relative recovery of mutagenic and carcinogenic activity by these short term assays. The reverse osmosis procedure as currently used also yields residue amounts adequate for in vivo toxicity testing. It is hoped that comparisons of toxicity data from in vitro and in vivo assays will contribute to the assessment of the health effects of mixtures of organic compounds in drinking water.

ACKNOWLEDGEMENTS

A portion of this study was supported by research grant R804202 from the U.S. En-vironmental Protection Agency to the University of Cincinnati.

REFERENCES

1. F.C. Kopfler, W.E. Coleman, R.G. Melton, R.G. Tardiff, S.C. Lynch, and J.K. Smith, Extraction and identification of organic micro pollutants - reverse osmosis method, Annals N.Y. Academy of Sciences, in press (1977).

2. R.G. Tardiff, Carcinogens in the aquatic environment: Identification and toxicological studies of organic chemicals, in Proceedings Fourth Annual Car-cinogenesis Collaborative Conference, Orlando, Fl (1976).

3. B.N. Ames, J. McCann, and E. Yamasaki, Methods for detecting carcinogens and mutagens with the Salmonella/mammalian microsome mutagenicity test, Mutation Res. 31, 347 (1975J:

4. T. Kakunaga, A quantitative system for assay of malignant transformation by chemical carcinogens using a clone derived from Balb/3T3, Int. J. Cancer 12, 463 (1973).

5. I. MacPherson and L. Montagnier, Agar suspension culture for the selective assay of cells transformed by polyoma virus, Virology 23, 291 (1964).


6. V.F. Simmon, K. Kanhanen, and R.G. Tardiff, Mutagenic activity of chemicals identified in drinking water, in the International Conference on Environmental Mutagens, to be published by Elsevier Scientific Publishing Co. (1977).

7. G.A. Junk, C D . Chriswell, R.C. Chang, L.D. Kissinger, J.J. Richard, J.S. Fritz, and H.J. Svec, Applications of resins for extracting organic components from water, 2., Anal. Chem. 282, 331 (1976).

8. E. Yamasaki and B.N. Ames, Concentration of mutagens from urine by adsorption with the nonpolar resin XAD-2: Cigarette smokers have mutagenic urine, Proc. Nat. Acad. Sei., U.S.A. 74, 3555 (1977).

Biomedical Aspects of Biorefractories in Water

H. F. KRAYBILL, Ph.D* C. TUCKER HELMES, Ph.D. and CAROLINE C. SIGMAN, Ph.D.**

^National Cancer Institute, Bethesda, Maryland **Stanford Research Institute, Menlo Park, California

I. INTRODUCTION

It has been postulated that a high percentage of human cancers have their origin in the environment and thus could be avoided if preventive measures and control procedures were instituted. Consequently, sophisti-cation of analytical procedures and biological methods for measurement of effects have been underway relative to micro-pollutants in the environment. In this area of concern, the modern instrumentation being used can detect parts per billion and parts per trillion levels of volatile organics in raw and potable water. Many reports have appeared in the literature on the detection, identification and classification of carcinogens (1,2,3,4). One of the most comprehensive conferences covering identification, biological effects, and epidemiological aspects of water biorefractories was that held in New York City in September 1976, sponsored by The New York Academy of Sciences, entitled "Aquatic Pollutants and Biological Effects with Emphasis on Neoplasia" (5).

Under the power of the Safe Drinking Water Act of 1974 (PL 93-523), the Environmental Protection Agency requested the National Academy of Sciences to conduct a study of the adverse health effects attributable to the carcinogenic and/or noncarcinogenic contaminants in drinking water. This report, issued in 1977, gave attention to all potential contaminants in water, chemical and biological (6). Recognizing that there are several million chemicals in the universe (3), those of natural origin and those from industrial technologies, it is evident that the pollution problem becomes a difficult one to control. Newer methods and approaches are needed for assessment of the carcinogenic potential of a whole class of chemicals and mixtures of chemicals in the water, atmosphere, diet and the workplace. These conceptual approaches are described elsewhere by Kraybill (7).

419

420 H. F. Kraybill, C. Tucker Helmes and C. C. Sigman

II. ORIGIN OF CARCINOGENS IN WATER

Even if it were technologically feasible to obtain a liter of water completely free of organic and inorganic contaminants, the cost would be prohibitive. Raw water supplies, away from industrial and population centers where pollution would seem escapable, may contain cations and anions of geologic origin. Some organic molecules are transported through the atmosphere and through fallout are dispersed into lakes and streams. The process is reversed in that chemicals such as DDT, a suspect carcinogen, are volatilized from bodies of water and transported through the upper atmosphere and are ultimately deposited in other regions of the world. Attesting to such a prevailing mechanism is the fact that the polar icecaps of Antartica and the fat of penguins contain DDT. This shows the extent of migration of DDT (8).

Natural seepage problems of oil and polycyclic aromatic hydrocarbons (PAH) into the continental shelf and beaches pose a real public health problem. Wilson and coworkers (9) have estimated the release of carcino-genic PAH into the marine environment to be at the rate of 0.2 X 10 to 6 X 10 metric tons per year. Forty percent of the world's total seepage from the ocean floor occurs in the Pacific Ocean; the Southern California coastal area being one of the seepage prone areas.

There are few waterways that do not carry inorganic material such as the heavy metals and asbestiform materials. Rivers flowing through areas near asbestos mines have higher concentrations of asbestos fibers. For example, in the Marin County area in California, fiber counts approached 2 X 10 fibers per liter (10). However, Andrew (11) reported that 31 streams in Minnesota and Wisconsin had no detectable concentration of amphiboles. Nicholson et al (12) conducted analyses on some eastern river waters in the USA and found chrysotile in the micrograms per liter range. The range in average values was from 1.3 to 5.9 micrograms per liter.

The concern about asbestos fibers in drinking water, foods, and parenterals (asbestos filters used) evolved from prior knowledge that inhaled asbestos induced mesotheliomas in occupational exposures. Since no definitive data was available on the potential hazard of ingested asbestos, a conference was held in North Carolina by the National Institute of Environmental Health Sciences and the Environmental Protection Agency in 1973 to address this problem (13). This concern was accelerated by the observation in the Duluth, Minnesota (Lake Superior area) where an iron ore company had been dumping, over the years, taconite or tailings of cummingtonite - grumerite and amphibole, into the lake which supplies the water for adjacent communities. Large amounts of amphibole appeared (31% of total solids), as reported by Cook et al (14) in 1974.

There is some evidence for the carcinogenicity of inorganic contaminants from both animal testing (15) and human observations. Berg and Burbank (16) con-ducted a study, oriented around the occurrence of trace metals in water supplies,

Biomedical Aspects of Biorefractories in Water 421

to determine cancer mortality rates in certain river basins in the USA. The product of the frequency of detection and the average detected concentration of the metal was calculated. This summary figure was used to determine the rank order of the basin for eight metals. These values are shown in Table 1. Berg and Burbank (16) found some correlations between levels of carcinogenic metals and cancer mortalities in certain regions. For example, nickel concentrations appeared to correlate with mouth and intestinal cancer death rates and arsenic with eye, larynx and myeloid leukemia. In Formosa and Argentina, arsenic in the drinking water has been implicated as a cancer cause. Beryllium was correlated with bone, breast, and uterine cancer. Lead was associated with kidney, stomach, intestinal and ovarian cancer, and leukemias and lymphomas.

There is some caution in assessment of the biological effects of these metal cations in drawing conclusions on statistical associations. The valence state of metal is important, as for example, hexavalent chromium is considered toxic at high levels and carcinogenic as an occupational exposure whereas trivalent chromium appears physiologically essential as the chromium insulin complex and glucose tolerance factor (19). The physical form and route of administration are important. Some metals and metalloids such as arsenic and selenium appear to be required metabolically at low levels, thus implying a threshold (20).

The levels of metals in drinking water in the USA and the average intake per day of some of these trace elements are given in Tables 2 and 3 (21,22,23,24).

In the treatment of water for drinking purposes, the chlorination process -long considered an essential procedure for public health safety to destroy pathogens - concurrently converts certain molecular species into other organic carcinogenic molecules. Some organic carcinogens appearing in raw water either appear for the first time in drinking water or are escalated in concentration. Such is the case for chloroform which increases from a level of a few micrograms per liter in raw water to several hundred micrograms per liter in chlorinated drinking water. Similarly, some brominated compounds are formed through chlorination. The enhancement is believed due to the effect of chlorination on humic acids.

Carcinogenic organic pollutants are likely to enter wells, lakes, streams and estuaries, and ultimately the sea, by the following mechanisms:

(a) through effluents from manufacturing plants

(b) through municipal sewage effluents

(c) through erosion or run-off from forests and lands as agricultural chemicals or soil chemicals

(d) through acid mine drainage

(e) through deliberate spillage on ground reaching aquifers and wells

(f) through deliberate dumping from ships and untreated sewage barged to sea


TABLE 1 RANGES OF METAL CONCENTRATIONS IN TEN RIVER BASINS

METAL

ARSENIC BERYLLIUM CADMIUM CHROMIUM COBALT IRON LEAD NICKEL

LOW FREQUENCY DETECTION

2 0 0 5 1 79 12 2

BASIN MEAN OF POSITIVE LEVEL

(Jo) ^JG/LITER

53

BERG AND BURBANK, REFERENCE 16.

IABLE

--17 1

13 8 5

2 TOXIC TRACE METALS

HIGH FREQUENCY OF DETECTION (%)

9 14 8 56 10 99 24 25

IN MUNICIPAL WATERS 100 LARGEST UNITED STATES CITIES

METAL MICROGRAMS/L

ESSSENTIAL

CHROMIUM

VANADIUM

MANGANESE

IRON

COBALT

COPPER

ZINC

MOLYBDENUM

FLUORIDE

35 70

1,100 1,700

0.01 250

2.1 68 7

BASIN MEAN POSITIVE LEVEL JJG/LITER

68 0.3 3 14 19 120 33 31

METAL MICROGRAMS/L

TQXIC

CADMIUM

MERCURY

NICKEL

LEAD

90 -

34 62

SLIGHTLY TOXIC

BARIUM

TIN

380 0.005

NQN TQXIC

ALUMINUM 1,500 BORON 590 STRONTIUM 1,200

TITANIUM

LITHIUM

49 170

MODIFICATION OF TABLE ON DATA ON WATER, DURFOR AND BECKER (1964), REFERENCE 22.


TABLE 5 AVERAGE INTAKE OF TRACE METALS IN DRINKING WATER

AVERAGE CONC. MICROGRAM INTAKE % OF SAMPLES W I T H

METAL MICROGRAMS/LITER AT 2 LITERS/DAY MICROGRAM/LITER OR MORE

CADMIUM

CHROMIUM

COBALT

COPPER

IRON

LEAD

MANGANESE

NICKEL

SILVER

ZINC

1.3 2.3 2.2

134.5

166.5

13.1

22.2

4.8

0.8

193.8

2.6 4.6

4.4

269.0

333.0

26.2

44.4

9.6

1.6 387.6

63 11

62

99

99

74

78

78

23

100

MODIFICATION OF DATA FROM ANGINO, WIXSON AND SMITH (1977), REFERENCE 24.

(g) through natural seepage from the ocean floor

(h) through the transfer process from atmosphere to water and the reverse process with recyclic deposition in areas quite distant from the transfer process.

III. IDENTIFICATION, QUANTIFICATION, DISTRIBUTION AND CLASSIFICATION OF ORGANIC CARCINOGENS, COCARCINOGENS AND PROMOTERS IN WATER SUPPLIES

Previous reports by Kraybill (2,3,4) and the National Academy of Sciences (6) have provided an identification, quantification, distribution and classification of organic carcinogens in drinking water. These listings were based on information obtained in 1975 and 1976. In this report we have extended our search to roughly 1700 unique organic chemicals to include 309 biorefractories in drinking water (25,26,27). A comprehensive listing of 1259 different compounds that have been found in one or more of 33 different water types (28) and a compilation of polluting substances found in fresh water, effluent discharges in aquatic animals and sediments are also listed from a European study (29). A systematic classification of the carcinogenicity of these chemicals has not yet been completed since for many of them new carcinogenicity bioassay results are not available at this time. For others a surveillance of the remaining 1500 or more chemicals needs to be achieved through bioassay or other screening procedures to classify their carcinogenic activity.


The objectives of the task referenced in this report were to evaluate the available data on the carcinogenicity of the chemicals identified in water by the aforementioned surveys and to classify these chemicals as either recognized carcinogens, suspected carcinogens, cocarcinogens or promoters, chemicals with inadequate test results, and chemicals not tested or of unknown carcinogenicity.

Where possible, the assessment for carcinogenicity was based on a review of available compilations of evaluated carcinogenicity data such as the IARC (International Agency for Research on Cancer) Monographs (30) and the American Chemical Society Monograph 173 on Chemical Carcinogens (31). This review was supplemented by an evaluation of data from the exten-sive compendium of the Public Health Service Publication No. 149 (32), and other sources such as the National Library of Medicine (Bethesda, Maryland) Toxline/Cancerline data base.

A. The Classification Process

(1) Recognized Carcinogens - The following criteria are used in classifying a chemical as a recognized carcinogen: (a) chemical established as human carcinogen by epidemiolo-gical studies, (b) in experimental studies chemical causes a statistically significant increase in the incidence of tumors or a decrease in latency period for the induction of tumors compared to controls, or demonstrates a dose-response effect in the induction of tumors, and is confirmed as being active in more than one species, or, by two or more different laboratories, or, (c) chemical appears on lists, generally accepted by scientists, of chemical agents that have demonstrated carcinogenicity in man or experimental animals. The recognized carcinogens are listed in Table 4.

(2) Suspected Carcinogens - The evidence for the carcino-genicity of these chemicals is not conclusive. Although they have been found to be carcinogenic in limited studies, the results either were not statistically significant, were not confirmed in more than one species or in several strains of one species, or in more than one laboratory even if in the latter cases the results were slightly significant for a species or strain of animal. The suspect carcinogens are listed in Table 5.

(3) Tumor Promoters or Cocarcinogens - Several of the chemicals that have been found in water have been reported to be active as tumor promoters or cocarcinogens. These are listed in Table 6.

(4) Chemicals of Unknown Carcinogenicity - Most of the chemicals that have been found in water are of unknown carcinogenicity. These either have not yet been tested adequately for carcinogenicity in biologically meaningful studies or have been tested and found to be inactive within the limits of these experiments.


TABLE 4 RECOGNIZED CARCINOGENS (RAW AND DRINKING WATER)

425

Name

Acrylonitrile CAS No. 107131

4-Aminob ipheny1 CAS No. 92671

4-Amino s t ilb ene CAS No. 834242

Benzene CAS No. 71432

Benz(a)anthracene CAS No. 56553

Benzo(a)pyrene CAS No. 50328

Benzo(b)fluoranthene

__ Species

Rodent, man

Mouse, rat, man

Rat, mouse

Man

Mouse

Mouse, rat, hamster, rabbit, monkey

Mouse

Site

Ear canal, mammary, intestine, lung

Bladder, liver, mammary

Acoustic duct, bladder, mammary

Leukemia

Liver, lung, skin, subcutaneous

Skin, lung, fore-stomach, subcutane-ous, mammary

Skin, subcutaneous

References

33, 34, (27,28,29)

35

36, (29)

37, 38, (26,27,28)

39, (28,29)

40, (26,27, 28,29)

41, (28,29) CAS No. 205992

Benzidine CAS No. 92875

Bis(chloromethyl)ether CAS No. 542881

Carbon tetrachloride CAS No. 56235

Chloroform CAS No. 67663

Dibenz(a,h)anthracene CAS No. 53703

1,4-Dioxane CAS No. 123911

Ethyl carbamate CAS No. 51796

Ethylene dibromide CAS No. 106934

Rat, mouse, dog, man Bladder, liver 4^, (28,29)

Mouse, rat, man Lung, nasal cavity, 4_3, (28) skin, subcutaneous

Mouse, hamster, rat Liver

Mouse, rat Liver, kidney

Mouse, rat

44, (26,27, 28,29)

45, 46, (26, 27,28,29)

Forestomach, skin, 4Λ (28,29) lung, subcutaneous

Rat, guinea pig Liver, nasal cavity, 48, gall bladder (26,27,28)

Mouse, rat, hamster Lung, liver, skin 4£, (28,29) lymph

Mouse, rat Stomach 50, 51, (28)


TABLE 4 (CONTINUED) RECOGNIZED CARCINOGENS (RAW AND DRINKING WATER)

Name

Ethylenethiourea CAS No. 96457

Indeno(1,2,3-c,d)pyrene CAS No. 193395

2-Naphthylamine CAS No. 91598

Species

Mouse, rat

Mouse

Site

Liver, lymph, thyroid

References

52, (3)

Skin, subcutaneous .53 > (28,29)

Mouse, dog, hamster Bladder, liver J54·, (28) man

4-Nitrobiphenyl CAS No. 92933

o-Toluidine CAS No. 95534

Safrole CAS No. 94597

Vinyl chloride CAS No. 75014

Notes: 1) References

Dog, man

Mouse, rat

Mouse, rat

Mouse, rat, man

underlined relate to

Bladder

Liver, bladder, subcutaneous, reticuloendothelial

Liver

Liver, lung, mammary

55, 56, (28,29)

57, 58, (28,29)

59, (29)

60, (26,27,28)

evaluation and classification of carcinogenic activity. References in parentheses relate to sources which report on these chemicals in raw and drinking water.

2) Listing relates to those chemicals on which evidence (experimental and epidemiological) was available as of September 1977. As more bioassay and human data become available the list will be expanded.

3) Classification as recognized is in some cases judgemental based on criteria used in this paper.

Biomedical Aspects of Biorefractories in Water 42.7

TABLE 5 SUSPECT CARCINOGENS IN WATER (RAW AND DRINKING WATER)

Name Species Site References

Aldrin CAS No. 309002

Mouse Liver 61, (26, 27,28,29)

Azobenzene Mouse CAS No. 103333

Benzo(j)fluoranthene Mouse CAS No. 205823

Benzo(k)fluoranthene Mouse CAS No. 207089

Benzo(g,h,i)perylene Mouse CAS No. 191242

BHC, Technical grades Mouse CAS No. 608731

Bis(2-chloroethyl)ether Mouse CAS No. 111444

2-Bromoethylpropane Mouse CAS No. 78773

Butylbromide Mouse CAS No. 26602891

Liver ^ 2 , (29)

Skin 63, (28,29)

Skin, subcutaneous 64,65, (29)

Skin

Liver

Liver, skin, subcutaneous

Lung

Lung

66,67, (29)

68> (29)

69, (26, 27,28,29)

70, (28)

70, (26, 27,28,29)

Chlordane CAS No. 57749

Mouse Liver 71, (26, 27,28,29)

Chloromethyl methyl ether Mouse, rat, man(?) Lung, skin CAS No. 107302 subcutaneous

72, (3)

Chrysene CAS No. 218019

Mouse Skin 71, 72., 21, 74,(28, 29)

Dieldrin CAS No. 60571

Mouse Liver 75, (26, 27,28,29)

DDE CAS No. 72559

Mouse Liver 76, (26, 27,28,29)

DDT Mouse Liver CAS No. 50293

76, (26, 27,28,29)


TABLE 5 (CONTINUED) SUSPECT CARCINOGENS IN WATER (RAW & DRINKING WATER)

Name

2,4-Dimethylaniline CAS No. 95681

1,2-Diphenylhydrazine CAS No. 122667

Ep ichlorohydrin CAS No. 106898

Heptachlor CAS No. 76448

Heptachlor epoxide CAS No. 1024573

Kepone CAS No. 143500

Lindane CAS No. 58899

Methyl iodide CAS No. 74884

Mi rex CAS No. 2385855

Species

Rat

Mouse, rat

Mouse

Mouse

Mouse

Mouse, rat

Mouse

Mouse, rat

Mouse, rat

Nitrilotriacetic acid (NTA) Mouse, rat CAS No. 139139

Nitropropane Rat CAS No. 79469

Pentachloronitrobenzene Mouse CAS No. 82688

Polychlorinated biphenyls Mouse CAS No. 1336363

Propylene oxide CAS No. 73569

Quinoline CAS No. 91225

Rhodamine B CAS No. 81889

Saccharin CAS No. 81072

Simazine CAS No. 122349

Rat

Rat

Rat

Rat

Rat

Site

Liver, subcutaneous

Liver, mammary, urinary tract, skin

Subcutaneous

Liver

Liver

Liver

Liver

Lung, subcutaneous

Liver

Urinary tract

Liver

Liver, skin

Liver

Subcutaneous

Liver

Subcutaneous

Bladder

Subcutaneous

References

77, (28)

113,(26,27)

78, (28)

79, (26, 27,28, 29)

80, 81, (26, 27,28,29)

113,(26,27)

J52, (27,28)

7(), 83., (27, 28)

84, 85, (28)

86, 87., (29)

j$8, (28)

113, (26,27)

89., 90, (28, 29)

91, (28)

92, (28,29)

.93, (28,29)

94, (28)

95, 96, (27,28,29)


TABLE 5 (CONTINUED) SUSPECT CARCINOGENS IN WATER (RAW AND DRINKING WATER)

Name

1,1,2-Trichloroethylene CAS No. 79016

1,1,2-Trichloroethane CAS No. 79005

1,1,2,2,-Tetrachloroethane CAS No. 79345

1,1,2,2-Tetrachloro-ethylene CAS No. 127184

Tannic acid (Tannins) CAS No. 1401554

2,4,5-Trichlorophenoxy-acetic acid CAS No. 93765

Toxaphene CAS No. 8061352

p-Toluidine CAS No. 106490

Vinylidene chloride CAS No. 75354

Species

Mouse

Mouse

Mouse

Mouse

Rat

Mouse

Rat, mouse

Mouse

Mouse

Site

Liver

Liver

Liver

Liver

Liver

Lung, liver, leukemia

Liver

Liver

Kidney

References

iZ., (27, 28, 29)

98, (26, 27,28,29)

99, (28,29)

100, (26, 27,28,29)

101,

102,

103, (28,

105,

106, (26,

(29)

(28)

104, 29)

(28)

27, 28)

Notes:

1) References underlined relate to evaluation and classification of carcinogenic activity. References in parentheses relate to sources which report on these chemicals in raw and drinking water.

2) Listing relates to those chemicals on which evidence (experimental and epidemiological) was available as of September 1977. As more bioassay and human data become available the list will be expanded.

3) Classification as recognized is in some cases judgemental based on criteria used in this paper.

430 H. F. Kraybill, C, Tucker Helmes and C. C, Sigman

TABLE 6 CHEMICALS WITH TUMOR PROMOTING OR COCARCINOGENIC ACTIVITY

CAS No.

95487

95578

95658

95874

95954

105679

106489

108430

108689

108952

Name of Chemical

Ortho-Cresol

2-Chlorophenol

3,4-Xylenol

2,5-Xylenol

2,4,5-Trichlorophenol

2,4-Dimethylphenol

4-Chlorophenol

3-Chlorophenol

3,5-Xylenol

Phenol

112301

112403

112629

112801

112958

1-Decanol

n-Dodecane

Methyl oleate

Oleic acid

Eicosane

120809

120832

123013

124185

138863

Pyrocatechol

2,4-Dichlorophenol

Dodecyl benzene

n-Decane

Limonene

143077

192972

334485

593453

629594

Dodecanoic acid

Benzo(e)pyrene

Decanoic acid

Octadecane

n-Tetradecane

1120214

8046535

n-Undecane

Alkyl benzene sulfonate

Total Promoters/Cocarcinogens » 27


The initial phase of this survey for carcinogenicity was directed toward assigning unequivocal identification to each of the chemicals and eliminating all duplications. This was accomplished, where possible, by assigning a Chemical Abstracts Services Registry Number (CAS No.) to each chemical. For approximately 270 of the chemicals it was not possible to assign CAS numbers. Hence, unequivocal identification of these was not possible and data retrieval was complicated. Some of the chemicals, for example, were listed as unspecified isomers. Where possible, CAS numbers were found for specific isomers and the classification of carcino-genicity was based on data for these. As an illustration, in Table 4, 4-nitrobiphenyl, 4-aminobiphenyl and 2-naphthylamine were listed in the sources as unspecified isomers.

A summary of the results of this survey is presented in Table 7. Since there is a potential for human exposure to these chemicals and such a large number of them were judged to be inadequately tested or found not to have been tested at all, there is clearly a need for more bioassay research to determine the carcinogenicity of these chemicals.

In addition to the listing of contaminants in the water supplies of the United States, the listing compiled by the Commission of the European Communities was used, making a total listing of 1728 chemicals for this survey on carcinogenicity (29).

TABLE 7 SUMMARY OF SURVEY RESULTS

CLASSIFICATION FOR NUMBER OF CARCINOGENICITY CHEMICALS

RECOGNIZED CARCINOGENS 22

SUSPECTED CARCINOGENS 42

TUMOR PROMOTERS/COCARCINOGENS 27

CHEMICALS - INADEQUATE TEST RESULTS 314

CHEMICALS NOT TESTED 1323

TOTAL FOR ALL CLASSES 1728

IV. IDENTIFICATION OF MUTAGENS IN WATER SUPPLIES

To complement any study on carcinogenicity, mutagenicity data on chemicals is useful, especially for those of unknown carcinogenicity. Chemicals found to be mutagens could be considered potential carcinogens and thus such indicative data would ascribe a high priority to testing chemical(s) for carcinogenicity.

To extend the study on carcinogenicity of water chemicals, mutagenicity data on these chemicals has been evaluated and presented in tabular form. For this study the evaluation of mutagenicity was based on tests of the chemical in histidine-requiring strains of Salmonella typhimurium developed by Ames (107) . Other references al-so provide a background on

431


the correlation between a chemical's mutagenic activity in these bacterial strains and its carcinogenicity (108,109).

We do not propose to go into any detail on these tests since such details are provided in a report by Sigman et al (110). For data and information from the literature, the Environmental Mutagen Information Center (EMIC) data base was used and such material was furnished by Shelby (111). Also, recent issues of prominent journals on toxicology, cancer research, mutation research, and environmental contamination were utilized (110).

In order to be consistent with the previous evaluation of the contaminants in water (raw and drinking), the results on the mutagenicity evaluation are presented in categories of chemicals that are "recognized," "suspect," and of "unknown carcinogenicity." Similarly, within these categorical designa-tions of carcinogenic activities, the results on mutagenic activity are recorded as mutagens or suspected mutagens. The criteria for classification of a mutagen or suspected mutagen are those set forth by Sigman et al (110). A chemical was classified as a "mutagen" if it was demonstrated to cause a dose-related increase in revertants in the Ames1 test. The minimum number of revertants recognized as evidence of activity was twice the spontaneous rate at one or more of the doses tested. Chemicals with activity demonstrated at only a single dose and those reported to be active, but for which no supporting evidence was available, were classified as "suspected mutagens." Additional testing or more detailed reporting of the experimental data were considered to be necessary for evaluating the mutagenicity of these chemicals. Only those with positive mutagenicity data are recorded in the following tables. A fuller account on the evaluation of all 1728 chemicals, including reports on negative tests for mutagenicity, are described elsewhere by Sigman et al (110). As was done previously, the chemicals are listed alphabetically and, therefore, the Chemical Abstract Numbers (CAS No.) are not necessarily in ascending numerical order.

For the 65 chemicals identified as mutagens or suspect mutagens, some duplication appears in the source material for test data acquired from the literature survey and the specific test data of Simmon et al (112). In a study conducted by Stanford Research Institute (SRI) for the Environ-mental Protection Agency (EPA), 27 chemicals were identified as mutagens; additionally, 46 were identified as mutagens from a literature survey. There were 8 duplications in information in the recording of literature-survey data and the recent EPA/SRI laboratory survey data on some 71 chemicals.

The tabular material presented will identify these two sources of data. For example, in the literature search, 34 mutagens and 13 suspect mutagens were identified. From the EPA/SRI study there were 23 mutagens and 3 suspect mutagens. Again, subtracting the 8 duplications in test data from both sources, the total number of chemicals identified as mutagens or suspect mutagens was 65.


In Table 8 data is presented on those water chemicals which are recognized carcinogens and identified as mutagens or suspected mutagens. Table 9 presents data on those water chemicals which are suspected carcinogens and which are also classified as mutagens or suspected mutagens. In Table 10 data is given on water chemicals of unknown carcinogenicity and subsequently classified as mutagens or suspected mutagens. The mutagenic activity is identified by the Ames1 procedures. The referenced material for each chemical encompassing 65 mutagens comprises about 150 specific references and are included in the report of Sigman et al (110). A recapitulation of this information and data is presented in a summary in Table 11 of available mutagenicity data (Ames1 procedures) on 1728 chemicals identified in water.

Excluding promoters and cocarcinogens in Table 7 there were 61 chemicals listed with some carcinogenic activity. Thus, there are 1667 from the listing of 1728 on which there is no carcinogenicity data. Among this group, however, 27 chemicals tested by the Ames1 Salmonella test systems were demonstrated to be mutagens. Using the Ames* systems, an additional 9 chemicals were classified as suspected mutagens from limited tests for activity with substantiating experimental data.

Thus, this class of chemicals for which no carcinogenic data is available contains some with mutagenic activity suggesting the possibility of potential carcinogenicity. This possibility is strenghtened by the observation that 78 percent (14/18) of the water chemicals which are recognized carcinogens and 60 percent (15/25) of those that are suspected carcinogens were found to be mutagens or suspected mutagens by the Ames1

procedure. These mutagens of unknown carcinogenicity would be prime candi-dates for carcinogenicity bioassay research.

It is obvious that a large number of chemicals identified in water (raw and potable, sediments and discharges) have not been tested for carcino-genicity (1667) or mutagenicity (1562). While man may not be directly exposed to all of these in the form of drinking water, marine animals and other biota prevail in these environments and thus many of these chemicals become part of the food chain as aquatic contaminants. Additionally, many of these on which there is non-existent data may also occur elsewhere in the environment as air pollutants and food contaminants and, thus, impose an additional environmental stress. For these multiple reasons it is imperative to identify any other potential carcinogens from this broad group of chemicals. To achieve these goals, a structure activity analysis of the chemicals of unknown carcinogenicity is planned by Stanford Research Institute.

V. ESTABLISHMENT OF RISK FACTORS

The potential hazard of ingesting aquatic pollutants or biorefractories can be assessed by two approaches: one by a series of epidemiological studies and the other by traditional toxicity studies in laboratory animals from which risk factors may be derived. The merits and limitations of the statistical/epidemiological studies will be discussed later.

Unfortunately, in the animal studies, appraisement is invariably made from extrapolations relevant to high doses, orders of magnitude beyond the usual concentration, and on single agents. Occurrence in the aquatic

TABLE 8 WATER CHEMICALS WHICH ARE RECOGNIZED CARCINOGENS: MUTAGENS AND SUSPECTED MUTAGENS

MUTAGENS

CAS No.

75014

NAME

107131

92671

834242

56553

50328

92875

53703

106934

96457

91598

92933

Acrylonitrile(a) (a)

4-Aminob ipheny 1 (a)

4-Aminostilbene

(a)

1,2-Benzanthracene

■D

ί

\ (a,b)

Benzo(a)pyrene

Benzidine(a)

(a)

1,2,5,6-Dibenzanthracene

(a)

Ethylene dibromide

(a)

Ethylenethiourea (a)

2-Naphthylamine (a)

4-Nitrobiphenyl

Vinyl chloride (a,b)

S. TYPHIMURIUM

STRAINS IN WHICH

MUTAGENIC ACTIVITY

OBSERVED

TA1535, TA1978

+

(TA98)C, TA100, TA1537, TA1538 +

TA100, TA1538

+

TA98, TA100, (TA1537)°,

+

(TA1538)C

TA98, TAIOO, TA1537, TA1538

+

TA1538

+

TA98, TAIOO, TA1537, (TA1538)° +

TAIOO, TA1530, TA1535

G-46, TA1530, (TA1535)C, (G-46)C

TAIOO, TA1535, TA1538

TA98, TAIOO,

(TA1538)C

TA98, TAIOO, TA1530, TA1535

+

MAMMALIAN METABOLIC

ACTIVATION

Mouse

Phenobarbitol Pretreatment


Phenobarbitol and 3 MCA

Pretreatment



Aroclor 1254 or Phenobarbitol

Pretreatment

Phenobarbitol or

Aroclor 1254 Pretreatment

(with TA100)

A.

TABLE 8 (CONTINUED) WATER CHEMICALS WHICH ARE RECOGNIZED CARCINOGENS: MUTAGENS AND SUSPECTED MUTAGENS

B.

SUSPECTED MUTAGENS

CAS No.

542881

94597

NAME

(a)

Bis(chloromethyl)ether

Safrole(a)

S. TYPHIMURIUM

STRAINS IN WHICH

MUTAGENIC ACTIVITY

OBSERVED

G-46

TA1530, TA1532

MAMMALIAN METABOLIC

ACTIVATION

(a)

(b)

(c) Data from survey of literature - references 107, 108, 109, 110, 111.

Data courtesy of EPA/SRI, Tardiff, Simmon, Kauhanen - reference 112.

For chemicals classified as mutagens, strains enclosed in parentheses, e.g., (TA98), are ones in which

activity was observed, but no positive dose-response was demonstrated.

+

Activity in presence of S-9 microsome fraction from liver of rats unless species otherwise specified;

Animals pretreated with Aroclor 1254 to induce microsomal enzymes unless other inducing agents

specified.

- Activity in absence of S-9 microsome fraction.

+

Activity both in presence and absence of S-9 microsome fraction.

+

TABLE 9 WATER CHEMICALS WHICH ARE SUSPECTED CARCINOGENS:

MUTAGENS

MUTAGENS AND SUSPECTED MUTAGENS

CAS No.

103333

111444

109659

78762

507197

57749

218019

106898

74884

91225

75354

NAME

Azobenzene

Bis(2-chloroethyl)ether^ ^

n-Butyl bromide(a,b)

sec-Butyl bromide

'

tert-Butyl bromide

Chlordane(b)

(a)

Chrysene

(a)

Epichlorohydrin

Methyl iodide(a,b)

(a)

Quinoline

Vinylidene Chloride ^

5 >b)

S. TYPHIMURIUM

STRAINS IN WHICH

MUTAGENIC ACTIVITY

OBSERVED

TA100

TA100

TA100

TA100, TA1535

TA100

TA100

TA98, TA100

G-46, TA100

TA100

TA98, TA100

TA100

TA1530

MAMMALIAN METABOLIC

ACTIVATION

S3

P3

0

PL

+ Mouse; Phenobarbitone

Pretreatment

TA1535

+ Phenobarbitol

Pretreatment

A.

+ + + +

TABLE 9 (CONTINUED) WATER CHEMICALS WHICH ARE SUSPECTED CARCINOGENS: MUTAGENS AND SUSPECTED MUTAGENS

B.

SUSPECTED

CAS No.

191242

127184

79345

79016

MUTAGENS

NAME

(a)

Benzo(g,h,i,)perylene

(a)

Perchloroethylene

(a)

l,l,2,2-Tetrachloroethanev '

Trichloroethylene

'

S. TYPHIMURIUM

STRAINS IN WHICH

MUTAGENIC ACTIVITY

OBSERVED

TA98

TA100

TA1530, TA1535

TA100

TA1535, TA1538

MAMMALIAN METABOLIC

ACTIVATION

+

- -

+ Mouse, B6C3F1

-

(a)

(b) Data from survey of literature - references 107, 108, 109, 110, 111.


+

Activity in presence of S-9 microsome fraction from liver of rats unless species otherwise specified.

Animals pretreated with Aroclor 1254 to induce microsomal enzymes unless other inducing agents

specified.

-

Activity in absence of S-9 microsome fraction.

+

Activity both in presence and absence of S-9 microsome fraction.

TABLE 10 WATER CHEMICALS OF UNKNOWN CARCINOGENICITY;


A.

MUTAGENS

CAS No.

192972

108601

74975

75274

75252

74839

124481

2051629

107073

75296

107051

96139

3018120

107062

NAME

(a)

Benzo(e)pyrene

Bis(chloroisopropyl)ether

Bromochloromethane

Bromodichloromethane

Bromoform

Bromome thane

Chlorodibromomethane

(a)

4-Chlorobiphenylv

2-Chloroethanolv

J

2-Chloropropane

1-Chloropropene

(a)

2,3-Dibromopropanol

Dichloroacetonitrile

(a)

l,2-Dichloroethanev

'

S. TYPHIMURIUM

STRAINS IN WHICH

MUTAGENIC ACTIVITY

OBSERVED

TA100

TA100

TA100

TA100

TA100, TA1535

TA100

TA100

TA1538

TA100, TA1535

TA1530

TA100

TA100

TA100, TA1535

TA100

(TA100)C, TA1530, (TA1535)

TA1535

MAMMALIAN METABOLIC

ACTIVATION

+

+

+ Phenobarbitol Pretreatment

+ Mouse; Phenobarbitol

Pretreatment +

+

+ + +

TABLE 10 (CONTINUED) WATER CHEMICALS OF UNKNOWN CARCINOGENICITY:


A. MUTAGENS (CONTINUED)

CAS No.

78875

542756

606202

75218

74873

74953

75092

91623

612602

611325

108452

71556

118967

NAME

(a)

1,2-Dichloropropane (a

)

1,3-Dichloropropene

(eis and trans isomers)

2,6-Dinitrotoluene^ '

(a)

Ethylene oxidev

'

Methyl chloride(a,b)

Methylene bromide

Methylene chloride (a)

6-Methylquinoline (a

)

7-Methylquinoline (a

)

8-Methylquinoline

(a)

m-Phenylenediamine

1,1,1-Trichloroethane^

'

(a)

2,4,6-Trinitrotoluenev

J

S. TYPHIMURIUM

STRAINS IN WHICH

MUTAGENIC ACTIVITY

OBSERVED

TA100,

TA1535

TA1978

TA100, TA1535, TA1978

TA100

TA1535

TA100,

TA1535

TA100

TA100

TA98,

TA100

TA98,

TA100

TA98,

TA100

TA1538

TA100

TA98

MAMMALIAN METABOLIC

ACTIVATION

+

+

+

+

n>

&3

+ + +

TABLE 10 (CONTINUED) WATER CHEMICALS OF UNKNOWN CARCINOGENITICY; MUTAGENS AND SUSPECTED MUTAGENS

B.

SUSPECTED MUTAGENS

S. TYPHIMURIUM

CAS No.

112265

238846

63252

1300216

124403

121142

122145

87683

129000

NAME

1,2-Bis(chloroethoxy)eth

(a)

Benzo(a)fluorene

Carbaryl (Sevin)(a)

(a)

Dichloroethane

(unspecified isomer)

(a)

Dimethylamine

(a)

2,4-Dinitrotoluene

(a)

Fenitrothionv '

Hexachloro-1,3-butadiene

Pyrene(a)

(b)

ane

(b)

STRAINS IN WHICH

MUTAGENIC ACTIVITY

OBSERVED

TA100

TA98

TA1538

TA100

TA1530

TA1535

TA100

TA100

TA100

MAMMALIAN METABOLIC

ACTIVATION

+ Mouse; Phenobarbitol

Pretreatment

+ 1 +

H

C o

v

' Data for survey of literature - references 107, 108, 109, 110, 111.


(c) For chemicals classified as mutagens, strains enclosed in parentheses, e.g., (TA98), are ones in which

activity was observed, but no positive dose-response was demonstrated.

+

Activity in presence of S-9 microsome fraction from liver of rats unless species otherwise specified.

Animals pretreated with Aroclor 1254

to induce microsomal enzymes unless other inducing agents

specified.

-

Activity in absence of S-9 microsome fraction.

+_ Activity both in presence and absence of S-9 microsome fraction.

0

(c


TABLE 11 SUMMARY OF AVAILABLE MUTAGENICITY DATA

(AMEST PROCEDURES) ON 1728 CHEMICALS IDENTIFIED IN WATER

BY CARCINOGENICITY CLASS

A. RECOGNIZED (SEE TABLE 8)

1) MUTAGENS 12

2) SUSPECTED MUTAGENS 2

3) CHEMICALS TESTED FOR WHICH THERE IS NO EVIDENCE OF ACTIVITY 4

4) CHEMICALS FOR WHICH TEST DATA WERE NOT FOUND 4

SUBTOTAL 22

B. SUSPECTED (SEE TABLE 9)

1) MUTAGENS 11




SUBTOTAL 42

C. UNKNOWN (SEE TABLE 10)

1) MUTAGENS 27




SUBTOTAL 1665

GRAND TOTAL A + B + C 1728

TOTALS - ALL CLASSES

MUTAGENS 50

SUSPECTED MUTAGENS 15

CHEMICALS TESTED - NO EVIDENCE OF ACTIVITY 101

CHEMICALS - NO TEST DATA FOUND 1562

TOTAL 1728


environment is multiple. This suggests the potential for additive effects, enhancement (e.g., tumor promotion or cocarcinogenesis), interactions, and, perhaps, some inhibiting effects. Thus, at best, extrapolations are uncertain. They are suggestive but not necessarily irrefutable in terms of causality. Since in terms of a neoplastic response, low-level insults may require a long-term induction period in man and indeed in animals, lifetime studies with hundreds of thousands of rodents may be necessary to establish a significant rise in cancer incidence for a test group over a so-called non-test (control) group - if then. However, even observations from control groups are confounded by contaminants in the air, the diet and drinking water furnished. If an effect occurs from a low-level insult over a long period, by the time the induction period is achieved the carcinogenic effect may be irrever-sible. Unfortunately, early warning signs are not yet available to predict, in time, such occurrence of neoplastic lesions.

Quantitatively, the acute effects at higher doses or maximum tolerated doses may be similar in a spectrum of animals. Metabolically and bio-chemically, however, species and strain variation could be dependent on metabolic pathways, the genetic background for enzyme profiles involved in these pathways, or detoxification mechanisms. On a body weight basis, man may be more reactive than experimental animal models; perhaps by an average factor of ten. Additionally, on the basis of pharmaco-genetics, some members of the human population may be more highly suscep-tible. Thus, a safety factor must be built in to protect these groups in the population. While some of these pharmacological/toxicological approaches are suitable in non-carcinogenic events for arriving at safety predictions, they are usually less satisfactory in dealing with irreversible toxicology such as cancer and extrapolation to zones of low-level, long-term exposure where current knowledge, with data and information, is essentially unavailable.

Certain principles are considered in the evaluation of the carcinogenic risk which are covered in many textbooks; however four succinct points or principles are from a recent report released by the National Academy of Sciences - National Research Council, USA, Committee on Safe Drinking Water (113).

The above document discusses some salient points such as (a) properly qualified, effects in animals are applicable to man, (b) methods do not now exist to establish a threshold for long-term effects of toxic agents, (c) exposure of experimental animals to toxic agents in high doses is a necessary and valid method of discovering possible carcinogenic hazards in man, and (d) material should be assessed in terms of human risk rather than as safe or unsafe. These four conceptual items are guidelines. It is not possible to discuss each in detail in this paper as to lines of agreement, disagreement or exceptions, if you will, on these seemingly universally acceptable tenets. This subject is covered elsewhere in more detail (114). As laudable as some of these guiding principles may be as an operational basis for approaching safety decisions, the counterarguments raised by scientific groups, especially in areas of pharmacological response and predictions from high dose to low dose response, are really reflective of our current state of ignorance and lack of information and data on mechanisms to derive one's interpretation and decisions on carcinogenic hazard.


For the organic solutes (volatile and nonvolatile) which are dealt with in this paper, the National Academy of Sciences Committee (113) looked at approximately 300 volatile organic compounds and 55 pesticides identi-fied in drinking water. It should be emphasized that the organic compounds identified in drinking water make up only a small fraction of the total organic components. About 90 percent of the volatile organic compounds have been identified and quantified but, again, these represent only about 10 percent by weight of total organic material. For the non-volatile components that comprise this 90 percent fraction, only 5 to 10 percent have been identified. The volatiles are readily analyzed by gas liquid chromatography and mass spectrometry.

Some of the carcinogenic organic contaminants previously listed in Tables 4, 5, 8, 9 and 10 were examined by the National Academy of Sciences Committee in a risk assessment study. For these compounds, where lifetime animal feeding studies were available, a statistical extrapolation of risk was performed using appropriate mathematical models. The risk to man was expressed as the probability of cancer being induced

by the continuous daily exposure to these carcinogenic organic contaminants over a 70 year lifespan, assuming the daily ingestion of 1 liter of water which contained a measured amount, i.e., 1 microgram per liter for these organic contaminants. Thus, estimates were derived based on the actual concentrations in the water. The statistical and mathematical procedures for arriving at these risk factors are described in the National Academy of Sciences Committee Report. In Table 12 the various carcinogenic contaminants and their highest concentrations in finished water and the statistically derived estimates of risks are presented based on the previous assumptions of lifetime intakes. The numbers in Table 12 are upper 95 percent confidence estimates of risk to man from a lifetime exposure to a single chemical. Nothing is stated about the integrated effect of all carcinogens. Corrections have been made for interspecies variations (i.e., from rodent to man) on the basis of relative surface area.

From examination of this data and any other derived data using mathematical models and statistical approaches the very low levels of environmental con-taminants, that is those in water, are on a statistically derived basis capable of inducing a finite number of tumors or cancers in a national popula-tion of 100 million people. It must be remembered that many assumptions are built into these derivations such as shape of the dose-response curve and slope of the curve. Whether biologically such events occur in the shallow portion of the dose-response curve, applicable to low level intakes in water, remains to be established. In essence, these mathematical extrapolations may be quite conservative and indeed overstate the risk. Perhaps, if one includes the possibility of additive insults, the reported risk factors for the single agents may be within the zone of actual risk for the multiple integrated insult for all carcinogens.

The terminology, "animal carcinogen" and "human carcinogen," used in the original table from the National Academy of Sciences Committee Report (113) on which Table 12 was based, was not used since the term "animal carcinogen" implies that the chemical is not a human carcinogen. The latter could ulti-mately be the case for chemicals tagged as human carcinogens.

The classification in Table 12 is not consistent with National Academy of Sciences Committee classification in other respects but is consistent with the classification in this paper. For example, recent epidemiological studies on benzene as a leukemogen in pliofilm workers would establish this chemical as a recognized carcinogen.


TABLE 12 ESTIMATION OF SOME LIFETIME CANCER RISK FACTORS (MAN) FROM

EXPERIMENTAL ANIMAL DATA AND KNOWN CONCENTRATIONS OF

RECOGNIZED AND SUSPECTED ORGANIC CARCINOGENS IDENTIFIED IN DRINKING WATER

UPPER 95% CONFIDENCE ESTIMATE OF HUMAN LIFE-TIME CANCER RISK PER MICROGRAMS/LITER * CAS NO. NAME

A. RECOGNIZEDJCAJLCINOGENS

71432

50328

51235

67663

96457

75014

BENZENE^

BENZO(A)PYRENE^

CARBON TETRACHLORIDE^b)

CHLOROFORM ^

ETHYLENETHIOUREA b '

VINYL CHLORIDE^

B. SUSPECTED CARCINOGENS

309002

111444

57749

60571

50293 72559

122667

76448

ALDRIN

BIS(2-CHLOROETHYL)ETHER

CHLORDANE

DIELDRIN

DDT/DDE

DIPHENYLHYDRAZINE

HEPTACHLOR

MAXIMUM OBSERVED CONCENTRATION IN FINISHED WATER MICROGRAMS/LITER

10 , (one value at 300)'

DETECTED3

5

366

NOT DETECTED

10

DETECTED3

0.42

0.1

8

DETECTED3

1

DETECTED3

1024573 HEPTACHLOR EPOXIDE

143500 KEPONE

58899 LINDANE

ALPHA BHC

BETA BHC

DETECTED

NOT DETECTED

0.01

DETECTED3

DETECTED3

I.D.

I.D.

1.5 X 10

3.7 X 10

2.2 X 10

5.1 X 10 -7

I.D.

1.2 X 10

1.8 X 10

2.6 X 10~

1.2 X 10

I.D.1

4.2 X 10

I.D.1

4.4 X 10"

9.3 X 10"

6.5 X 10

4.2 X 10

-6

-5

-5

-5

-6

-6


TABLE 12 (CONTINUED) ESTIMATION OF SOME LIFETIME CANCER RISK FACTORS (MAN) FROM

EXPERIMENTAL ANIMAL DATA AND KNOWN CONCENTRATIONS OF

RECOGNIZED AND SUSPECTED ORGANIC CARCINOGENS IDENTIFIED IN DRINKING WATER

CAS NO. NAME

MAXIMUM OBSERVED CONCENTRATION IN FINISHED WATER MICROGRAMS/LITER

UPPER 95% CONFIDENCE ESTIMATE OF HUMAN LIFE-TIME CANCER RISK PER MICROGRAM/LITER*

1336363 PCB (AROCLOR 1260)

82688 PCNB

122349 SIMAZINE

79016 TRICHLOROETHYLENE

NOT DETECTED

< 0.1

0.5

(Water, Milan, Italy - up to 450)' 4

(one value at 13,000)

3.1 X 10

1.4 X 10 -7

6.4 X 10 -8

1.3 X 10 -7

Modification of Data from NAS Committee Report, Reference 113.

In authors' view, Atrazine (1912249), Endrin (72208) and Propazine (139402) are of unknown carcinogenicity or negative.

Notes:

(a)

(b)

1,

2

Epidemiological evidence convincing as recognized carcinogen.

Animal evidence strong to classify as recognized carcinogen.

Insufficient data to make extrapolations.

In Water - Federation of American Societies for Experimental Biology/Depart-ment of Defense Report on Evaluation of the Health Aspects of Certain Compounds Found in Irradiated Beef, Bethesda, Maryland, August, 1977.

Detected but not quantified.

Data from Connecticut State Department of Health, 1977.

See text for explanation on derivation of risk factors.

7


VI. EPIDEMIOLOGICAL CONSIDERATIONS AND ASSESSMENT OF EPIDEMIOLOGICAL STUDIES

There have been many statistical studies on contaminants in water and their probable associations with cancer mortalities. Most of these, reported at various meetings, have not been reported in the literature. The index of credibility for some of these statistically derived associa-tions of aquatic contaminants with cancer mortalities is not very high. One study that might fall in this area is that by Burton and Cornhill (115). These investigators based a study on 100 of the largest USA cities (population combined over 60 million) where certain quality criteria such as specific conductivity, total dissolved solids, and hardness, appeared to be correlated with cancer mortalities (age adjusted) in these cities and counties where cancer mortality data was available. Several reviewers of this report have indicated that this study is highly tentative and until the analysis is refined in several respects and the flaws in analysis corrected, the report, as written, adds little to our knowledge and perhaps raises more questions than it provides answers.

Another study focused on the parishes of Louisiana because of suspicions of the possible etiologic role of drinking water in the high incidence of bladder cancer in New Orleans (116). This relationship would appear to hold for cancer of the urinary organs and gastrointestinal tract. There are some caveats to be attached to this study in that the effect of occupational exposures could be buried in the aggregate data. Further-more, cancer rates in New Orleans could be inflated by migration of people from other industrial areas and exposures, as well as regional effects of diet and the possible contribution of air pollutants in a metropolitan area. Data on tobacco and alcohol exposures and air pollutants by parishes were unavailable. As is true in other statistical associations, drinking water contaminants in causality of cancer in this study is not established but provides a basis for further exploration.

Other studies in the Ohio River valley were conducted. Reiches (117) reported that Ohio citizens who drank water from the Ohio River and Lake Erie have a cancer mortality rate 8 percent higher than those Ohions who drink Scioto River water. She contends that this study incriminates chloroform as a possible cause. At this moment, in the absence of a penetrating review of the data, it is difficult to ascertain why this chemical was singled out when there are other potent carcinogens that appear in water, all of which could contribute singly, or in an additive way, to the total carcinogenic insult. An unpublished report by Buncher (118) proposes that cancer incidence in the vicinity of Cincinnati - a population ingesting ground water - differs from that reported for those who obtain their drinking water from surface water. This report has not been critically evaluated.

Cantor and associates (119) approached the problem of studying the potential etiology of water contaminants by adopting a descriptive approach to generate a theory. They examined the potential role of the non-chloroform trihalomethanes (NCTHM). These trihalomethanes such as bromoform, dibromo-chloromethane and dichlorobromomethane have been shown to have mutagenic activity (see Tables 9 and 10) but, experimentally, their carcinogenic activity has not been determined in rodents. These chemicals are planned for bioassay by the National Cancer Institute. In the aforementioned study,


the period 1968 to 1971 was chosen for the cancer mortality data in man. Water exposure data was derived from the USA National Organics Reconnaisance Survey of 1975 and a special survey conducted by the Environmental Protection Agency on Region V. For specific correlation purposes, data on 76 counties in the USA, where they had adequate water supply data and where a monitored water supply was furnished to more than 50 percent of the population, were utilized.

The reason NCTHM was used is that this trilogy of halogenated compounds appeared to show a stronger association than chloroform with certain cancer mortality rates, especially bladder cancer in men and women. This latter point is worth emphasizing when referencing the earlier hypothesis by Reiches (117) on the etiologic contribution of chloroform. The NCTHM factor appears to have some association with bladder cancer mortality. Although other sites were correlated, bladder cancer appeared to be a common factor and certainly more highly correlated, positively, than, for example, brain cancer for those counties having more than 85 percent of the population served by the monitored water supply. By moving those counties that appear to have high relative values for NCTHM and cancer mortalities by at least three standard deviations away from the predicted values, the above conclusions still seem to hold. Geo-graphically, bladder cancers in females seem to be the same in all regions of the USA. For brain cancers there are better correlations in the North and West portions of the USA.

The authors of this report are quick to point out that the consistency of these experiences should be given more weight than any specific significant statistical sssociations. As indicated previously, cancer of the bladder seems to be the common osbervation which requires or suggests further investigations. These preliminary observations seem to be consistent with other statistical studies reported in the USA. The caveat, as always, is that these preliminary studies are only suggestive and perhaps are good indicators for further studies buttressed by the corollary data and convincing evidence.

There are other unpublished reports on contractually supported studies by the Environmental Protection Agency at several universities that have as their primary objective the comparison of cancer rates in communities that chlorinate with those that do not. These studies are either still in the planning stage or currently too incomplete to produce any conclusive findings. Future reports, however, should elicit some findings as to whether the chlorination versus non-chlorination process has any potential influence in the induction of cancer (120).

One troublesome feature with some of these statistical associations made in the various studies is that correlation coefficients do not prove that two factors have a cause and effect relationship. They show only associations that may exist. This association could arise because both factors are related to other factors and hence appear related to each other. Stated otherwise, other variants could be contributing in parallel fashion (i.e., air pollutants, drugs, food contaminants, etc.). The correlation coefficients in some of the studies observed are exceedingly low. This assumes a large sample size to make such small correlation coefficients statistically significant.


VII. REINFORCEMENT OF STATISTICAL-EPIDEMIOLOGICAL STUDIES

Investigations in the area of micro exposures are tedious and complexed by other contributing factors of similar low level exposures that must be controlled in the derivation of cause and effect relationahips. The high level exposures to potential chemical carcinogens in the workplace are obviously more easily explored in terms of goals for establishment of cancer causation for exposed populations. Ancillary studies from the experimental area, or ecological and/or epizootic observations, may lend some support in the reinforcement of statistical studies oriented around population groups. Some of these approaches are worth emphasis.

Studies are currently underway at various laboratories to assess the carcinogenic and mutagenic activity of the array of chemicals that have been identified and quantified in water. These have been previously mentioned. Such data provide some suggestive evidence for further follow-up in human population studies. The work of Tardiff and colleagues (121) relating to studies on concentrated extractable residues and solvent fractions of these residues searching for activity - carcinogenic and mutagenic - in a certain class of contaminants is most instructive and, no doubt, rewarding. The details on preparation of the concentrate from 500 gallons of municipal water and testing in bacterial systems and cell cultures are described elsewhere (25).

Another procedure that could be adapted to screening of specific biore-fractories or a composite of biorefractories is a rapid assay method based on Chinese hamster lung fibroblasts (122). These fibroflasts are cultured with radioactive thymidine grown in a non-radioactive medium and finally exposed for one to four hours to the test substance in the presence of liver cell microsomes. These microsomes are added to simulate metabolic processes and hence render otherwise inactive agents carcinogenic by metabolic activation. The extent of damaged DNA by the test agent is ascertained by measuring the intact and filterable radioactive DNA after treated cells are disrupted and subjected to alkali.

The feasibility of using pre-screen mutagenic information in selective assays of water biorefractories for carcinogenic activity has been previously emphasized. These searches should uncover a host of new candidates and mixtures of components for further studies in the etiology of cancer.

Indirect and direct approaches have been used to characterize carcinogenic contaminants in water. The direct approach relates to identification, quantification and monitoring of their presence in water supplies where certain population groups receive these exposures. Cantor et al (119) used an approach based on the indicator, non-chloroform trihalogenated methanes. An indirect approach is to record the level and presence of chemical carcinogens in a water supply on the basis that such chemicals were previously demonstrated as carcinogens in animal models. Chloro-form is a good example of an indicator chemical but other recognized carcinogens could be utilized if occurring at significant levels and in great frequency.


One of the most intriguing possibilities is to use the marine animal as an experimental model for bioassay of single chemicals and a simulated mixture of chemicals as they appear as biorefractories in drinking water. This approach has one advantage in that observavations are based on a species that inhabits an environment where such exposures are relevant and continuous. Observations on tumorigenicity and incidence of various tumors could be correlated with chemical profiles on waters and fish tissues analyzed for body burden of chemicals encountered. Brown and coworkers (123) approached this procedure somewhat less elegantly, but perhaps more pragmatically, when they found that there was a higher frequency of tumors in fish in polluted waters of the Fox River watershed as compared to fish in relatively non-polluted waters. The incidence of tumors found in 2121 fish examined from the polluted watershed was 4.48 percent compared to a tumor incidence of 1.03 percent in 4369 fish taken from non-polluted waters. The fish bioassay also permits an opportunity to conduct comparative pathology between the fish model and mammalian model, the rodent, relevant to dose-response relation-ships in terms of tumorigenicity. Such an approach is in the planning stages at the National Cancer Institute.

The acquisition of such data in an ecological system is a forerunner to studies on epizootics where feral populations (fish) could be surveyed for tumor incidence in various surface waters or river systems that serve as water supplies for various municipalities. Profiles on the water, the sediments, and the fish, correlated with fish tumor incidence, could be the presumptive evidence for extension of such studies to human population laboratories. Such evidence could be used for explorations on human cancer mortality experiences in various geographical regions. It is reflective of a multidisciplined, interagency, programmatic approach that may reinforce the current statistical studies. It is also an approach seriously considered by the National Cancer Institute in some demonstration projects. The conceptual approaches suggested are represented schematically in Figure 1.

VIII. SUMMARY AND ASSESSMENT

It is evident that micro-pollutants in our water supplies, raw and finished, are a source of toxic agents, some of which are carcinogens and mutagens. The identification and quantification of a fraction of these potentially harmful agents is a tribute to modern science and technology. However, the fact that many remain to be identified and characterized as biologically active chemicals is not reassuring in terms of their possible long-term effects on marine and terrestrial animals and man. In addition to the requirement for qualitative and quantitative determination of many volatile organic components, there remains the task of a similar procedure on the non-volatiles. Also, the emphasis placed on the organic contaminants should be matched by a concerted effort on the inorganic contaminants which are no less biologically active in terms of adverse effects. Any carcinogenic insult to man could be the result of the integrated stress from organic biorefractories, inorganic contaminants, or a combination of these constituent chemicals.


The results of surveys of raw and treated water, and of contaminants in effluents and sediments have revealed a wide spectrum of recognized and suspected carcinogens as well as compounds of unknown carcinogenicity. At the present time, there are 22 recognized carcinogens, 42 suspected carcinogens, 27 chemicals with promoter or cocarcinogenic activity, 314 for which there are inadequate test results and the bulk, or remainder, of 1323 chemicals which have never been tested for carcinogenicity.

From this total group of 1728 surveyed for carcinogenicity and mutagenicity, 65 were classified with mutagenic activity (50 mutagens and 15 suspect mutagens). Twenty-seven of these 65 mutagens or suspected mutagens fall into the class of unknown carcinogenicity either because of no or inadequate tests. A first priority is then to match these 27 against any recent testing data. If no new data in uncovered, they would be good candidates for carcinogenicity bioassay, especially if any of this group of 27 were produced in large amounts or were exposure chemicals elsewhere in the environment (air, diet, workplace, etc.).

Since biological data, specifically dose-response data in the area of low-level exposures, is not available, risk assessments for the human population are made by extrapolating from experimental animal dose-response data on tumorigenicity to cancer incidence for a prescribed number for the population. Estimates derived from the application of various mathematical models provide calculated risk factors in terms of probable cancer cases for 100 million, 1 million, or other selected population size. While these risk factors are suggestive of potential risk for levels of exposure in the microgram per liter range, they are based on certain assumptions which may or may not hold relative to actual pharmacokinetic, pharmacological, or biochemical mechanisms which may prevail in a mammalian response to an environmental insult. Until a clearer biological picture is available as to events which can occur in the shallow portion of the dose-response curve, one is left to deal with such mathematical extrapolations. Whether the mammalian organism response to a high level insult in the milligram per kilogram of body weight or diet basis is the same as for an intake or stress of microgram or fraction of a microgram per liter exposure remains to be resolved pharmacologically and biochemically. There are illustrative cases which would indicate that the effects on the cells of the liver may not be similar and the metabolic pathways may be different under conditions of ä metabolic overload compared to a lower concentration stress. Perhaps the most significant, but often overlooked factor, is the integrated effect of all carcinogens and mutagens which additively could provide total exposures in the milligram per liter range. While standards and tolerances are developed for a single contaminant, it is conceivable that they may be reduced to provide for a certain margin of safety when multiple stresses are considered.

The series of epidemiological studies conducted relevant to site, sex, and race-specific cancer mortality rates, correlated with levels of drinking water contaminants in the United States, can serve as a working hypothesis towards more definitive analytical studies and perhaps provide some guidance in the regulation of these contaminants. Cancer mortality rates must be corrected for variants other than the water contaminant exposures. To the degree that it is possible to correct for other influences, such as socio-economic status, degree of urbanization and


manufacturing activity - all of which have a contributing effect - such derived associations will be dependent on how well such influences can be corrected. For establishment of a high index of credibility one needs reliable data on the other environmental exposure variables of interest, as well as information on other possible confounding variables, to permit for their correction in the statistical analysis. Improvement in the present generation of correlation studies is therefore dependent on the acquisition of a data base on a more extensive range of municipal supplies. Measures of monitoring other than those now in existence may be requisite to serve as useful and more accurate predictors of human toxicity where long-term, low-level exposures to aquatic contaminants (organic and inorganic) are concerned. Reservations on the applicability of such statistical associations in terms of cancer mortalities are commented upon when some epidemiologists advance the view that human studies on the carcinogenic effects of long-term, low-level exposures to drinking water contaminants are most difficult to perform so that the effort yields information which has some applicability in the etiology of cancer.

Recognizing some of the deficiences in these statistical studies, it seems appropriate to develop ancillary studies in the experimental area, using marine animal models to acquire data on tumor response that may be correlated with environmental profiles on chemicals that reflect qualita-tively and quantitatively the extent of exposures. Epizootics can make a significant contribution in support of the statistical human effects studies. Such an approach is proposed for the acquisition of suggestive evidence from the aquatic environment as a forerunner of human population studies or concurrently a resource for augmentation of such human studies in the quest for statistical associations that have great relevance with respect to causation. The achievement of such goals on causality relevant to long-term, low-level exposures requires the input from a multidiscipline approach reflecting project participation and collaboration of many govern-mental agencies, industrial groups, and international organizations.

Acknowledgements

The authors would like to acknowledge the assistance and cooperation in collating some of the information from the Environmental Protection Agency (Dr. R. G. Tardiff); the Environmental Mutagen Information Center (Dr. Michael D. Shelby) and the Stanford Research Institute staff (Dr. V. F. Simmon, L. Cheney, P. C. Hall, B. A. Lewin, M. A. Power, P. A. Sullivan and J. M. Walker). Appreciation is also expressed to Dr. E. K. Weisburger, National Cancer Institute for advice on classifi-cation of some of the carcinogens.

A. EXPOSURE TYPES

1.

2.

INGESTION

DRINKING H20 BIOREFRACTORIES

FISH RESIDUES - PROFILES ON CONTAMINANTS (BODY BURDEN)

EXPOSURE MONITORING - DOSE ESTIMATION

MUTAGENICITY ASSAY

BACTERIAL SYSTEMS

CELL CULTURES

DRINKING H20 BIOREFRACTORIES

a)

SINGLE CHEMICALS

b) MULTIPLE CHEMICALS

1

1

1 1

> IIES

W

RESPONSE

STATISTICAL ASSOCIATIONS

NEOPLASIA - MAN

1

<

1

1

PRESUMPTIVE AND ASSOCIATED EVIDENCE - EPIZOOTICS

(TUMOR INCIDENCE^- AQUATIC ANIMALS)

RODENT BIOASSAY

DRINKING HO BIOREFRACTORIES

a) SINGLE CHEMICALS

b) MULTIPLE CHEMICALS

EÖ

^

!*

σ1

1-»

H-1

O

C o S3

I-'

B

CO

P3 0 α

FERAL POPULATIONS

IDENTIFICATION - QUANTIFICATION

CONTAMINANTS IN RAW H

20

B.

FISH AS EXPERIMENTAL ANIMALS - CARCINOGENICITY BIOASSAY

1) SINGLE CHEMICALS

2) MULTIPLE CHEMICALS

Fig. 1. RELATIONSHIPS - ENVIRONMENTAL CANCER (AQUATIC ENVIRONMENT)

C/3

A.

B.


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34. J. M. Norris, Status report on the 2 year study in corporating acrylonitrile in the drinking water of rats. Special report from the DOW Chemical Company, Midland, Michigan, January 12, 1977.

35. International Agency for Research on Cancer, Evaluation of Carcinogenic risk of chemicals to man, Monograph 1, 74, WHO, Lyon, France (1972).

36. D. B. Clayson and R. C. Garner, American Chemical Society, Chemical Carcinogens Monograph 173, 395, Am. Chem. Soc. Publ., Washington, D. C. (1976)

37. International Agency for Research on Cancer, Evaluation of Carcinogenic risk of chemicals to man, Monograph, 7, 203, WHO, Lyon, France (1974).

38. P. F. Infante, R. A. Rinsky, J. K. Wagoner and R. J. Young, Leukemia among workers exposed to benzene (in press - The Lancet) (1977).


40. Ibid, 3, 91, (1973).

41. Ibid, 3, 69, (1973).

42. Ibid, 1, 80, (1972).

43. Ibid, 4, 231, (1974).

44. Ibid, 1, 53, (1972).

45. Ibid, 1, 61, (1972).

46. Department of Health Education and Welfare, Public Health Service, National Institutes of Health, National Cancer Institute, Report on carcinogenesis bioassay of chloroform, March 1 (1976).

47. International Agency for Research on Cancer, Evaluation of Carcinogenic risk of chemicals to man, Monograph 3, 178, WHO, Lyon, France, (1973).

48. Ibid, 11, 247, (1976).

49. Ibid, 7, 111, (1974).

50. W. A. Olson, R. T. Haberman, E. K. Weisburger, J. M. Ward and J. H. Weisburger, Induction of stomach cancer in rats and mice by halogenated aliphatic fumigants. J. Natl. Cancer Inst. 51, 1993 (1973).

51. M. B. Powers, R. W. Voelker, N. P. Page, E. K. Weisburger and H. F. Kraybill, Carcinogenicity of ethylene dibromide (EDB) and l,2-bibromo-3-chloropropane (DBCP). Tox. Appl. Pharm. 33 (1), 171, Abst. 123 (1975).



53. Ibid, 3, 229, (1973).

54. Ibid, 4, 97, (1974).

55. Ibid, 4, 113, (1974).

56. D. B. Clayson, Occupational bladder cancer, Prev. Med. 5 (2), 228, (1976).

57. A. B. Russfield, F. Homburger, E. K. Weisburger and J. H. Weisburger, Further studies on carcinogenicity of environmental chemicals including simple aromatic amines. Tox. and Appl. Pharm. 25, 446 (Abst. 20), (1973).

58. D. B. Clayson and R. C. Garner, Am. Chemical Society Chemical Carcinogens Monograph 173, Am. Chem Soc. Publ., Washington, D. C. p. 387. (1976).

59. International Agency for Research on Cancer, Evaluation of Carcinogenic risk of chemicals to man, Monograph, 10, 231, WHO, Lyon, France, (1976).

60. Ib id , 7, 291, (1974).

61 . I b id , 5, 25, (1974).

62. I b id , 8, 75, (1975).

63. I b id , 3, 82, (1973).

64. Public Health Service Report 149, items C-520, D-719, (1961-67 and Sec. 2, 1961-67), see reference #32.

65. A. Dipple, Am. Chemical Society Chemical Carcinogens Monograph 173, p.273, Am. Chem. Soc. Publ, Washington, D. C. (1976).

66. Public Health Service Report 149, items E-460, F-359, (1968-69; 1970-71), see reference #32.

67. A. Dipple, same as reference #65, p. 255.


69. Ibid, 9, 117, (1975).

70. L. Poirier, G. D. Stoner and M. B. Shimkin. Bioassay of alkyl halides and nucleotide base analogs in pulmonary tumor response on strain A mice. Can. Res. 35 (6), 1411, (1975).

71. Department of Health, Education and Welfare, Public Health Service, National Cancer Institute, Report on Carcinogenesis Bioassay of Chlordane, June 1, (1977).

72. International Agency for Research on Cancer, Evaluation of Carcinogenic risk of chemicals to man, Monograph 4, 239, (1974).

73. Ibid, 3, 159, (1973).

74. A. Dipple, same as reference #65, p. 245.


75. International Agency for Research on Cancer, Evaluation of Carcinogenic risk of chemicals to man, Monograph 5, 125, (1974).

76. Ibid, 5, 83, (1974).

77. Public Health Service Report 149, item G-260, (1972-73), same reference as 32.


79. Federal Register 39(229) 41299, Pesticide Products containing heptachlor or chlordane - Intent to Cancer Registration. Office of the Federal Register, Nat-ional Archives and Records Service, General Services Administration, Washington, D. C , November 26, 1974.

80. S. S. Epstein, Carcinogenicity of heptachlor and chlordane, Sei.total Environ. 6 (2), 103, (1976).

81. Editorial. EPA suspends use of heptachlor, chlordane, Chem. & Eng. News, 53 (31), 7, (1975).


83. P. D. Lawley, American Chemical Society Chemical Carcinogens Monograph 173, p 176, Am. Chem. Soc. Publ, Washington, D. C. (1976).

84. International AGency for Research on Cancer, Evaluation of Carcinogenic risk of chemicals to man, Monograph 5, 203, WHO, Lyon, France, (1974).

85. B. M. Ulland, N. P. Page, R. A. Squire, E. K. Weisburger and Cypher, R. L., A carcinogenicity assay of mirex in Charles River CD rats. J. Natl. Cancer Inst. 58 (1), 133, (1977).

86. DHEW, PHS, National Cancer Institute Carcinogenesis Technical Report #6, Nitrilotriacetic Acid, June 1, 1977.

87. Article, The Cancer Letter, 3 (22), 1, (1977) - Two more chemicals reported out of bioassay.

88. Report from the National Institute of Occupational Safety and Health, Chemical Engineering News, May 2, 1977, p. 10.

89. R. D. Kimbrough, R. A. Squire, R. E. Linder, J. D. Strandberg, R. J. Montali and V. W. Burce, Induction of liver tumors in Sherman strain female rats by polychlorinated biphenyl. J. Natl. Cancer Inst. 55 (6), 1463, (1975).


91. Ibid, 11, 191, (1976).

92. K. Hirao, Y. Shinohara, H. Tsuda, S. Fukushima and M. Takahashi, Carcinogenic activity of quinoline on rat liver. Cancer Res. 36, 329, (1976).

93. Public Health Service Report 149, item B-923, C-944, (1969, 1961-67), same reference as 32.


94. W. C. Lepkowski, Saccharin ban goes beyond issue of cancer. Chem. and Eng. News, 55 (15), 17, April 11, 1977.

95. K. Kay, A brief overview of the toxicological and epidemiological background to the detection and prevention of cancer in agricultural workers, Cancer Detec-tion and Prevention, 1 (1), 107, (1976).

96. R. Preussman, Chemische Karcinogen in der Menschlichen Umwelt, Handbook AIlg. Pathol. 6, 421, (1975).

97. DHEW, PHS, National Cancer Institute, Report on Carcinogenesis Bioassay of Trichloroethylene, Technical Report Series No. 2, February 1976.

98. Ibid, Report on Carcinogenesis Bioassay of 1,1,2-Trichloroethane, June 1, 1977.

99. Ibid, Report on Carcinogenesis Bioassay of 1,1,2,2-Tetrachoroethane, June 1, 1977.

100. Ibid, Report on Carcinogenesis Bioassay of 1,1,2,2-Tetrachloroethylene, June 1, 1977.


102. I. Muranyi-Kovacs, G. Rudali and J. Imbert, Bioassay of 2,4,5-Trichloro-phenoxyacetic acid for carcinogenicity in mice. Brit J. Cancer, 33, 626, (1976).

103. Article, EPA has begun a public inquiry on toxaphene safety, Chem. and Eng. News, May 30, 1977, p. 7.

104. Article, Toxaphene RPAR notice about to be issued by EPA has four "counts," Pesticide and Toxic Chemical News, May 18, 1977, p. 16.

105. Public Health Service Report 149, item G-295, (1972-73), same reference as 32.

106. C. Murray, Vinylidine chloride, no trace of cancer at Dow. Chem. Eng. News, March 14, 1977, p. 21.

107. B. N. Ames, Methods for detecting carcinogens and mutagens with Salmonella/ mammalian microsome mutagenicity test, Mutat. Res. 31, 347, (1975).

108. J. McCann, E. Choi, E. Yamasaki and B. N. Ames, Detection of carcinogens as mutagens in the Salmonella/microsome test: Assay of 300 chemicals, Proc National Academy Science, 72, 5135, (1975).

109. I. F. Purchase, E. Longstaff, J. Ashby, J. A. Styles, D. Anderson, P. A. Lesevre and S. R. Westwood, Evaluation of six short-term tests for detecting organic chemical carcinogens and recommendations for their use, Nature 264, 624, (1976).

110. C. C. Sigman, J. M. Walker and P. A. Sullivan, The mutagenicity of chemicals identified in waters (Ames' procedure). Report to the National Cancer Institute by Stanford Research International, Menlo Park, California, 94025, June 29, 1977.


111. M. D. Shelby, Environmental Mutagen Information Center Data Base, Oak Ridge, Tennessee, USA, May, 1977.

112. V. F. Simmon, K. Kauhanen and R. G. Tardiff, Mutagenicity activity of chemicals identified in drinking water, unpublished report of the Environmental Protection Agency and Stanford Research Institute (1977).

113. National Academy of Sciences/National Research Council, Assembly of Life Sciences, Advisory Center on Toxicology, Committee on Safe Drinking Water, Summary Report: Drinking Water and Health, Publ. by Nat. Acad. of Sciences, Washington, D. C. (1977).

114. H. F. Kraybill, From mice to men: Predictability of observations in experimental systems - their significance in man. Proc. of Conf. on Human Epidemiology and Animal Laboratory Correlations in Chemical Carcinogenesis, June 1-4, 1977, Mescalero, New Mexico, USA, (in press).

115. A. C. Burton and J. F. Cornhill. Correlation of cancer death rates with altitude and with the quality of water supply of 100 largest cities in the USA. (Manuscript furnished to author for review), (1977).

116. T. Page, R. H. Harris and S. S. Epstein, Drinking water and cancer mortality in Louisiana, Science 193, 55, (1976).

117. N. Reiche, Studies on cancer mortalities in populations exposed to water contaminants in Lake Erie, Scioto and Ohio Rivers, Toxic Material News, March 11, 1977, p. 99.

118. C. R. Buncher, Ohio study links cancer to contaminated drinking water. Toxic Material News, 3 (17), 133, (1975), August 18, 1976.

119. K. D. Cantor, R. Hoover, T. J. Mason and L. J. McCabe, Association of halomethanes in drinking water with cancer mortality, (in press, 1977).

120. L. J. McCabe, Personal communication to the authors from Health Effects Research Laboratory, Environmental Protection Agency, Cincinnati, Ohio (1977).

121. R. G. Tardiff and M. Deinzer, Toxicity of organic compounds in drinking water, Report presented at 15th Water Quality Conference, Feb. 7-8, 1973, University of Illinois, Urbana (unpublished, 1973).

122. Article, Mammalian cells are used in a new test that screens for chemical carcinogens, Chem. and Eng. News, 6, 18, Decl (1976).

123. I. Greenspan, J. B. G. Kwapinski and P. Beamer, Frequency of fish tumors found in polluted watershed as compared to nonpolluted Canadian waters. Cancer Res. 33, 189, (1973).

A Comparison of the Chlorinated Organic Compounds Present in the Fatty Surface Film of Water and the Water Phase Beneath ELIZABETH BAUMANN OFSTAD and GULBRAND LUNDE

Central Institute for Industrial Research, P. O. Box 350 Blindem, Oslo 3, Norway

The existence of a fatty surface layer of water has been well documented and this film has been shown to contain different pollutants like heavy metals, phtalates and chlorinated hydrocarbons (1-4).

A technique published by Larsson et al (3) for sampling of the surface film has been further explored in our laboratory. The technique is based on adsorbtion of the surface film to a perforated teflon sheet.

This sampling method has been used in an investigation of the persistent (sulphuric acid resistant) organic compounds present in the surface film. The analysis was performed by gas chromatography (EC-detector) and the results are obtained in gas chromatogrammes as "pollution profiles".

The samples have been obtained from different places in Norway including heavily waterpolluted areas as well as fairly uncontaminated areas. When sampling the sur-face film, the water underneath (0.5 m) were also collected and the "pollution profiles" of these water samples were similarly obtained by gas chromatography.

When the "pollution profiles" from the corresponding surface film and water samples were compared, marked differences were found between the two. Furthermore, all the surface film samples were found to contain substantial amounts of PCB. Com-pared to the water underneath the concentration of PCB in the surface film was estimated to be higher by a magnitude of 10 -10 . Other typical effluent pollu-tants present in some of the water samples were concentrated to a much smaller de-gree than PCB in the surface film (Fig. 1).

It is well known that PCB is transported in the environment by the air masses. The results obtained from this investigation indicates that a substantial part of the pollutants found in the surface film is deposited from the atmosphere.

Investigations of surface films of water by this technique has proved to be a useful tool in the work of characterizing air-transported pollutants in the envi-ronment. By comparing the "pollution profiles" of the surface film and water sam-ples, it may be possible to distinguish between air-deposited pollutants and pollu-tants originating from industrial or other effluents.

461

E. Baumann Ofstad and G. Lunde

SURFACE SAMPLE

230°C 8 °/min -

150°C

WATER SAMPLE

230°C 8 ° / m i n " 130°C Fig. 1. Pollution profiles of surface layer and

water masses from a heavily polluted area.

REFERENCES

(1) W. D. Garret, Deep-Sea Res. 14, 221-227 (1967).

(2) R. A. Duce, J. A. Quinn, C. A. Olney, S. R. Piotrowicz, B. J. Ray, T. L. Wade, Science 176 (14), 161-163 (1972).

(3) K. Larsson, G. Odham, A. S0dergren, Marine Chemistry 2, 49-57 (1974).

(4) W. D. Garret, J. D. Bultman, J. Colloid Sei. 18, 798-801 (1963).

462

Model Study of Lake Superior Organic Carbon Budget Gives Insight on Water Quality Trends

WALTERJ. MAIER* and WAYLAND R. SWAIN**

* University of Minnesota 55455 ^Environmental Protection Agency, Grosselle, Michigan 48138

Recognition that total organic carbon (TOC) is a sensitive, nonspecific indicator of water quality has prompted a study of the organic carbon budget of Lake Superior and its drainage basin with a view to quantifying the major sources and sinks of organic matter. The objective is to model the distribution of organics and to project the effects of major fluxes on the average organic carbon concen-tration (TOC) of the lake.

Influx of organic matter from watershed drainage averages-. 12.6 mg/1, ranging from 5 to 36 mg/1. This is equivalent to an influx of 6.3x10 grams of organic carbon per year compared to 1.4x10 g of TOC in the Lake.

Lake Superior Water Mass Watershed Drainage Rainfall Photosynthetic Production

Total (Yearly) Net (Yearly)

Concentration mg/1

1.1 12.6 3

0.13 0.04-0.01

.4xlo]^ g 1 6.3x10^ g/year 2.0x10 g/year

1.6xl0^9g/year 0.16x10^ g/year

Influx of carbon from the atmosphere comes down as rainfall and as dry fallout (dust). Since there are no published data on Lake Superior per se, atmospheric influx was deduced from data on regions with similar atmospheric conditions which show that rainfall contains from 0 to 3 mg/1 TOC. A value of 3 mg/1 is equivalent to an influx of 2x10 g/year of organic carbon. The net yearly accumulation of organic matter from photosynthesis is small for the lake as a whole although seasonal and localized concentration peaks may occur. Furthermore, in the course of a year, most of the synthesize biomass will be remineralized (75-90%) by the microbial flora. The net yearly addition to the TOC pool is estimated to range from 0.04 to 0.01 mg/1 for the lake as a whole.

Organic carbon is removed from the lake by direct outflow and by biochemical oxi-dation (mineralization) to form carbon dioxide. Outflow is proportional to the concentration of TOC in the lake water, which averages 1 mg/1, and the flowage rate which is 7x10 m /year and represents approximately 0.5% of the lake volume. Removal of organic matter by biochemical oxidation is probably the single most important factor that controls the overall carbon budget. Biochemical oxidation is known to occur in all surface waters; in polluted waters rates of oxidation are generally proportional to the concentration of organics. However, there is no pre-cise information on the rate of oxidation to be expected in the cold, unpolluted, low nutrient content waters of Lake Superior. The effects of biochemical oxida-

463

464 W. J. Maier and W. R. Swain

tion were therefore evaluated over a broad range of values deduced from studies on other surface waters.

A mathematical model that treats the lake as a well mixed basin was used to cal-culate the yearly average TOC levels of the lake. The effects of surface runoff, rainfall, photosynthesis, outflow and biochemical oxidation were evaluated separ-ately and in combination to assess long range effects. The results show that there is a large dilution volume that masks the effects of pollution inflows for many years. This effect is illustrated by the model projected TOC concentration in Figure 1. Using the most likely values for surface runoff, rainfall, and photosynthesis but no biochemical oxidation, the average TOC level of the lake is shown to increase substantially over the next 100 years. However, the effects would be difficult to detect for the first ten years. If biochemical oxidation rates were larger than 0.0005, TOC levels would decrease from present levels. This is a very unlikely situation and a more realistic assessment is that the rates of oxidation are lower and a gradual increase in TOC levels of Lake Superior will occur.

EFFECT OF BIOLOGICAL OXIDATION ON

ORGANIC CARBON CONCENTRATION

SURFACE RUNCFF COEFF!CIENT = 5g /m 2 - year

RAINFALL CARBON CONCENTRATION = 3mg/1

0.001

>.i I 1 1 I 1 10 100 1000

TIME, yr

Fig. 1. Effect of Biological Oxidation on Organic Carbon Concentration This relatively simple analysis of the Lake Superior carbon budget provides a useful framework for assessing the long term trends of TOC. The modeling studies show that relatively large quantities of organic matter enter the lake but their effect is not measurable on a year to year basis due to the large dilution volume which makes it difficult to detect increases in TOC levels which could herald large and irreversible long term changes in water quality.

Photochemical Reactions of 2,3,7,8- Tetrachlorodibenzo-p-dioxin (TCDD) Adsorbed on a Silica Gel Surface

HARUN PARLAR, SIEGMAR GAB and ISTVAN GEBEFÜGI

Institut für Ökologische Chemie der Gesellschaft für Strahlen- und Umweltforschung mbH München, D-8051 Attaching and Institut für Chemie der Technischen Universität München, D-8050 Freising- Weihenstephan

INTRODUCTION

To be able to estimate the influence of the chlorinated hydrocarbons on the quality of the environment, chemico-ecological and toxico-ecological data are primarily required. In addition to their level of production, their range of application and distribution tendency a knowledge of their transformation under environmental conditions is desirable. In contrast to the biological degradation pathways and transformation processes of environmental chemistry, whose course is accompanied by only very small amounts of energy, energy sources with an inexhaustible capacity and constant intensity are available for abiotic conversions. Temperature and UV-radiation become the most important forms of manifestation of the largest energy source of our environment, the sun, whose direct or indirect effect is especially important during an ecological assessment of chlorinated hydrocarbons. In this connection the task posed is an investigation of the reaction of chlorinated hydrocarbons under simulated environmental conditions. Such simulations are made more difficult by virtue of the complexity of these reactions. To be able to study individual reaction mechanisms, it is necessary to create conditions, which correspond to those in the environment, or make possible a partial simulation in order thus to enable the ex-perimental results to be interpreted with some degrees of proba-bility.

2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) is an important eco-logical compound because of its powerful toxic properties. Work carried out hitherto on the photochemical transformation of TCDD points to an only limited degradability in the presence of hydrogen donors such as oils, proton-containing solvents or pesticides. Under these conditions TCDD is converted into compounds with less chlorine while the carbon skeleton is preserved (2). No total photo-degradation of this compound has so far been reported. In the present

465

466 H. Parlar, S. Gab and I. Gebefugi

work an attempt is made to investigate the photochemical degrada-bility of TCDD adsorbed on a standardised carrier substance (Kieselgel 0.063 - 0.2 mm, Merck).

METHODS

An irradiation apparatus (V = 25 1) (3) was used for the UV irradia-tion. Varying the wavelengths was carried out with Pyrex ( Λ > 290 nm) or with quartz glass filters (A > 230 nm). A high pressure mercury lamp (HPK 125, Philips) served as light source. Respectively 240 /ig of the substance was drawn up approximately monomolecularly on 500 g silica gel using the method described by Gab et al (4). Both gas chromatography (Carlo-Erba Fraktovap 2200, 63Ni-EC, 300°C, 1.5m glass column, 3% SE 30 on Chromosorb W-AW-DMCS; oven temperature = 240°C) and a highly sensitive Bioassay-AHH induction method were employed to follow the course of the reaction.

RESULTS

From the results of the irradiation it was shown that TCDD adsorbed on silica gel is degraded relatively rapidly by the wavelengths above 290 nm (Pyrex filter) typical of the troposphere.

After irradiating for 7 days only 8% of the original TCDD was still detectable. The photochemical degradation is accelerated with the short-wave irradiation (quartz filter) and under these conditions only 2% residual TCDD is left after four days. These results show that 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) is relatively labile under our experimental conditions. Thus, with an adequate solar UV-irradiation level close to the ground photochemical de-gradation would be possible even without organic proton donors.

REFERENCES

(1) Accepted as Publication in part by Naturwissenschaften (1977).

(2) D. G. Crosby and A. S. Wong, Environmental degradation of 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD), Science 195, 1337 (1977).

(3) S. Gab, S. Nitz, F. Korte and H. Parlar, Aufbau und Funktion eines Reaktors für Reaktionen an festen Oberflächen, German Patent Application P 27 04 942.9 (1977).

(4) S. Gab, V. Saravanja and F. Korte, Irradiation studies of aldrin and chlordene adsorbed on silica gel surface, Bull. Environ. Contam. Toxicol. 13, 301 (1975).

Heavy Metals in Susquehanna River Bottom Sediments ~ Surficial Concentrations, Urban Impacts, and Transport Mechanisms

BRUCE McDUFFIE, IBRAHIM EL-BARBARY, GREGORY J. HOLLOD and ROBERT D. TIBERIO

Laboratory for Trace Methods & Environmental Analysis, Department of Chemistry, State University of New York at Binghamton, Binghamton, New York 13901, USA

Rivers, with characteristically variable flow regimes, have large variances in trace metal(TM) concentrations making it difficult to monitor pollution from urban or industrial areas. Thus, two stations upstream from Binghamton in 1974 yielded soluble Cd concentrations of 0.6 ± 0.4 yg/L (av. ± s.d., 6 samplings in duplicate per site), while two downstream sites gave 0.9 ± 0.2 ug/L, not significantly high-er. For other TM's, up- and downstream concentrations were even closer. In bottom sediments, however, distinctly higher values for Cd, Cr, Cu & Hg were found down-stream (Table 1), documenting the impact of the Binghamton area. Also, unexpect-edly high Ni and Zn levels in sediments at a rural site were found to be associated with upstream coal mine drainage.

Recent data on surficial HM's in 1976 silt samples from a 200-mile reach of the river (Fig. 1) show dramatically the impact of urban pollution from the Binghamton and to a lesser extent the Scranton-Wilkes Barre area, both with -200,000 people.

Heavy metal concentrations increase with decreasing particle size. The size-fraction ratios (Table 1) are more constant from site to site than the absolute concentrations, suggesting that cation exchange equilibria influence the distribu-tion. Also, element-element correlations among different samples, within a given size fraction, are strong only for a few pollutant metals (e.g., Ag & Cd, from a photographic industry), and there are few strong correlations with organic C.

TABLE 1 Heavy Metals in Susquehanna R. Basin near Binghamton, 1974-75.

River Water Bottom Sediments-Surficial Concentrations & Concentration Ratios Sol. Fraction* Cone., yg/g Ratios in Size Fractions

Ug/L Upstream Av. Downstream Av.

.0

(tot.) 0.018 0.082

Cd Cr Cu Hg Ni Pb Zn Fe Mn

0.7 1.8 5.8 0.024 5.0 5.8 6.9 --

0. 4. 12. 0. 16. 24. 52.

,000. 430.

6

018

3 11 24 0 21 21 66

10,000 304

Silt/Sand

4.5 1.8 3.4 3.9 2.5 1.7 1.9 1.5 1.6

Clay/Sand

10. 7. 14. 18. 8. 6. 5.0 4.0 2.1

* Av. of median values for 4 sites.

467

468 B. McDuffie, et al.

Thus organic matter may play less of a role than ion exchange in the uptake of TMfs by sediments.

Adsorption studies for Cd(II) and Pb(II) on reference clay samples and river sedi-ment, in a synthetic river water medium at pH 7, show the strong tendency of sedi-ment materials to take up trace metals and thus "buffer" the soluble TM levels in rivers. Such lab-loaded sediments release Cd completely in electrolyte media like sea-water and release both Cd and Pb completely in an acidic (pH 3.0) medium, sug-gesting the re-solubilization of TM's in estuaries or from acid discharges.

In the Upper Susquehanna Basin, sediment load varies approximately as the square of the water discharge, thus most transport occurs during storm events. Tropical Storm Eloise (Sep. 1975) caused a "cleansing" or scouring of the downstream sedi-ments for most pollutant metals (Fig. 2). This effect, and anomalous increases after the storm (organic C in sand, Fe and Mn in clay) warrant further study.

References: Reports (1975;1976) to the Susquehanna River Basin Commission, Mechan-icsbUrg, PA 17053 (whose support is acknowledged). See also Trace Substances in En-vironmental Health-X (1976), D.D. Hemphill, ed., Univ. of Missouri-Columbia, 85-95.

Fig. 1. Surficial heavy metals in Susquehanna River sediments. A. Map showing sampling sites. B. Concentrations in silt v§_ miles downstream.

Heavy Metals in Susquehanna River Bottom Sediments 469

diooo

100 150 RIVER MILES

o

CD

5 LU

> O z r-

> - I

< cc

O o

0.251

KEY: QSAND

[3 SILT

HCLAY Mn

Fig. 2. Impact of Tropical Storm Elolse: ratios of metal and organic C concentra-tions before and after the storm, in sand, silt, and clay fractions.

Nitrification in Rapidly Flowing Streams

W. GUJER

Swiss Federal Institute for Water Resources and Water Pollution Control, CH-8600 Dübendorf, Switzerland

The sequental biological oxidation of ammonium to nitrite and then to nitrate by Nitrosomonas and Nitrobacter respectively (nitrifica-tion) in shallow rivers (hydraulic radius <1 m) is mainly caused by sessile biomass. A mechanistic model for nitrification in biologi-cal films which may be applied to rapidly flowing streams (flow ve-locity^.5 m/sec) indicates temperature, heterotrophic plus photo-trophic biomass production, oxygen concentration and ammonium con-centration to be the parameters to influence this self-purification mechanism (Fig. 1) (1).

An ammonium mass balance for an entire river system, indicates that ammonium oxidation rates may be estimated with the proposed model, based entirely on published information (Table 1). This makes the model applicable to water pollution control planning.

Nitrite, as an intermediary product in the oxidation of ammonium to nitrate, is buffered by a biological equilibrium. In steady state the activity of Nitrobacter is compareable to the activity of Nitro-somonas . Based on bacterial growth kinetics (2), the equilibrium ni-trite concentration may be predicted in function of river water tem-

perature and bulk ammonium con-centration (Fig. 2) (3).Calculated and measured nitrite concentra-tions compare reasonably good, provided that ammonium and nit-rite concentrations are not in-fluenced significantly by domi-nant point sources in the imme-diate environment of the samp-ling station (Fig. 3) .

Because of the relationship be-tween bulk ammonium and bulk nitrite concentrations in rivers, it is indicated that the limita-tion of nitrite concentration, for reason of fish toxicity, must be supported by similar limita-

1 RIVER TIME

AKTIVITY

1 INFLUENT

CATCHMENT AREA

POPULATION

INDUSTRY

INFILTRATION

TOTAL INPUT

f NITRIFICATION Active Area

INCORPORATION Into Biomass

EFFLUENT

BALANCE (Input-Output)

G L A T T , Fällanden - Glattfelden

JULY 1 9 7 4 , Temperature = 20 °C

UNITS

4.6m3/sec

238 km2

231000lnh

I '340 000 m2

I000kg BM/d

8.7m3/sec

I LOAD C0FFFICIENTS NH4 - N

0.04 mg/I

50kg/km3-yr

3.5g/ Inh-d

-0 30g/m2-d

0.05 kgN/ kg BM

0.46 mg/1

N03-N

0.47 mg/l

, I500kg/km2-yr

5.0g/lnh-d

+-0 30g/m2-d

3.36 mg/l

I LOAD NH4-N

16

33

Θ09

--

858

-402

- 50

- 346

+ 60

(kg/d) 1 NO3-N

1 i8? 978 |

1155

-—

2 3 20 |

+ 402

-2526

+ 196

TABLE 1. Mass balance for inorga-nic nitrogen in the river Glatt(1)

471

472 W. Gujer

tions of the ammonium concentrations. At low pH values the nitrite tolerance limit may require lower ammonium concentrations than the limit for free ammonia.

Fallstudie Glatt",

References

1. W. Gujer, "Nitrifikation in Fliessgewässern · Schweiz. Z. Hydrologie, 38:2, 171-189 (1976).

2. G. Knowles, A.L. Downing and M.J. Barrett, "Determination of kinetic constants for nitrifying bacteria in mixed culture, with the aid of an electronic computer", J. Gen. Microbiology, 38, 263-278 (1965).

3. W. Gujer, to be published. 4. J. Zobrist, J.S. Davis and H.R. Hegi, "Charakterisierung des

chemischen Zustandes des Flusses Glatt", Gas-Wasser-Abwasser, 56, 97-114 (1976).

FIG. 1

Nitrification model for ra-pidly flowing rivers. The ni-trification rate is given as function of biomass production, temperature, and bulk NH cone. Indicated limits are 1) phototrophic production, 2) oxygen limitation, 3) wash-out of Nitrosomonas, 4) un-likely combination of biomass production and bulk NH4 cone, (from 1).

1 2 3 4 5

Biomass Production Rate

gBM/m 2 -day

^ 0.6 Γ T

-

-

-

~ Λ /

1 I . 1

RIVER WATER TEMPERATURE^

20 ° C / ^

/ 15*C^ "

/ ^ ^ JOJC^^^—

^ - ^ 5°C

, 1 1

-

-

_

-"

0.0 0.5 1.0 1.5 AMMONIUM IN THE RIVER (BULK CONCENTRATION)

mg NH4-N/I

TEMPERATURE RANGE 5-24°C BULK AMMONIUM CONC. 0.16-3.5

mgN/l

F I G .

Transferfunction model for ni-trite in rivers. Based on ni-trification kinetics in a bio-logical film, the equilibrium nitrite cone, is predicted from water temperature and bulk NH cone. (3).

FIG.

0.1 0.2 0.3 0.4 0.5

OBSERVED NITRITE CONCENTRATION mg NO2-N/I

3

0.6

Verification of transferfunc-tion model for nitrite. Zobrist et al (4) analysed 50 grab samples from the river Glatt in Switzerland (1974).

Fate of 3,3r-Dichlorobenzidine in the Aquatic Environment

H. APPLETON, S. BANERJEE, E. PACK andH. SIKKA

Syracuse Research Corporation, Syracuse, N.Y., U.S.A.

INTRODUCTION

Due to its widespread usage as an intermediate in the dye and pigment industry and its demonstrated carcinogenic properties, the discharge of 3,3f-dichlorobenzidine (DCB) into the aquatic environment is of concern to agencies charged with protection of human and environ-mental well-being. Due to the lack of information on the environ-mental fate of DCB, this study was initiated to determine the role of several mechanisms in governing the fate of DCB in the aquatic environment.

METHODS AND RESULTS

Sorption and Desorption of DCB by Aquatic Sediments

The degree of adsorption of DCB to natural aquatic sediments was determined by vigorous shaking of *^C-DCB-sediment suspensions (pH7, 20°C) in the dark. At appropriate time intervals, aliquots were centrifuged, and the degree of sorption determined radiometrically by disappearance of iI+C from the supernatant water. Equilibration was generally achieved within the first 24 hours of shaking. Dis-

tribution coefficients (u* ggg involution X 1 0 ° n l 8_1) for a number of aquatic sediments ranged from 2670 to 12800. Adsorption was in-hibited under alkaline (pH9) conditions by 30-50%. Desorption of DCB from sediment was very low, confirming the affinity of DCB for aquatic sediments.

Photodegradation of DCB

Irradiation of aqueous solutions of DCB with light from a 450W mer-cury arc filtered through Pyrex leads sequentially to mono-chloro-benzidine and benzidine, as well as to a number of highly colored water insoluble derivatives. The quantum yield for the disappear-ance of DCB exceeds unity, and the half-life for the degradation of DCB in natural sunlight is of the order of ninety seconds. However, in aprotic solvents such as cyclohexane or hexane, the chemical is

473

474 H. Appleton, et dl»

photostable for several hours. Photolysis of aqueous solutions of DCB or mono-chloro-benzidine at 2537A results in the formation of blue-green transients which decay in the dark. The rate of decay is of the first-order with respect to substrate, and is inhibited by acid. Possibly these transients are photosubstituted products which give rise to the colored insoluble materials.

Microbial Degradation of DCB in Lake Water

The degradation of DCB by naturally-occurring aquatic microbial communities was examined. 1I+C-DCB was incubated with lake water (21°C in the dark) obtained from a eutrophic lake and a mesotrophic lake. After a one month incubation period, the water samples showed 75% of original DCB when assayed by high performance liquid chroma-tography. No metabolites were detected. From these results, DCB appears to resist biodegradation by aquatic microbes.

Uptake, Elimination, and Metabolism of DCB in Fish

Bluegills (Lepomis macrochirus), 2"-3" in length were employed. Ex-posures were performed in darkness at 23°C under static conditions. When exposure times exceeded 24 hours, the fish were placed in fresh exposure baths at 24 hour intervals. Toxic levels of DCB were accumulated prior to establishment of a DCB equilibrium between water and fish. Fish exposed to 2.0 ppm 11+C-DCB showed whole body residues equivalent to 265 ppm DCB after 48 hours, while those exposed to 0.5 ppm contained 277 ppm after 120 hours. When residues exceeded 150 ppm, many mortalities ensued (the residues cited refer to sur-viving fish). Chemical concentrations were 1.5 to 3 fold higher in the offal than in the edible portion of the fish. Analysis showed the presence of DCB and an additional compound, which accounted for up to 60% of the total chemical burden. This material produced free DCB under mildly acidic conditions (pH2, 5 minutes). The behavior of the unknown is similar to that displayed by N-glucuronide or N-sulfate conjugates of aromatic amines. When exposed fish were placed in fresh, flowing water, 14C chemical rapidly decreased in the first day. However, elimination after this time was incomplete, with 11 to 13% of the ll+C-DCB and metabolite remaining both 5 and 10 days after initiating depuration.

CONCLUSIONS

The ability of DCB to concentrate in fish compounds the hazard invol-ved in release of DCB-containing effluents into water systems. Since DCB appears to resist microbial degradation, photodegradation and adsorption to suspended materials probably determine the persistence and bioavailability of DCB in the environment. However, photodegra-dation does not totally eliminate the hazard due to generation of benzidine, a known human carcinogen.

This investigation was supported by Grant R804-584-010 from the U.S. Environmental Protection Agency.

The Uptakej Tissue Distribution and Retention of Hexavalent Chromium by Young Rainbow Trout (Salmo Gairdnerii)

V. J. H. M. TEN HOLDER*, A. S. HOGENDOORN-ROOZEMOND*, Z. KOLAR**, J. J. T. W. A. STRIK * and J. H. KOEMAN*

*Department of Toxicology, Agricultural University, Wageningen **Interuniversity Reactor Institute, Delft, The Netherlands

With the aid of sodiumchromate solutions, containing 51Cr labelled chromate ions as radioactive tracer and measuring the radioactivity with a scintillation detector, the uptake and retention of Cr(VI) was followed in total alive fish as well as in the separate organs. Fish with an average weight of 0.2 g were used in the experiments, concerning the uptake and excretion kinetics and they were kept at a temperature of 13°C. The pH ranged from 7.8 to 8.0. In the experiments where the tissue distribution was studied, fish were used weighing 2 to 5 g. The chromium concentrations in the exposure solutions ranged from 1 to 10 ppm Cr. Uptake times varied from 1 to 9 days and excretion times were established from 1 to 20 days.

The uptake increased with increasing exposure concentration as well as with the exposure time. No definite conclusion could be drawn concerning the achievement of a steady state in chromium uptake.

Independent of the concentration and duration of exposure the kidney, liver and gills accounted for 3, 4 and 13% of the chromium absorbed resp.

After the fish, exposed to 1 to 5 ppm Cr, were transfered to Cr free water the excretion was followed by measuring the radioactivity and calculating the retention of chromium in the fish. It was observed that the chromium content in the fish decreased rapidly during the first few days and was relatively slow during the continuing period of excretion.

In order to separate the two phases in the excretion process, the results obtained were fitted to an equation consisting of a sum of exponential terms, by means of linear regression analysis and fcurve peeling1 (Ref. 1).

475

476 V. J. H. M. ten Holder, et dl.

General retention equation:

π, . -0.693 t/Tl' -0.693 t/Ti " R(t) = a- . e 2 + a . e 2

in which: R(t) = the fraction of the total accumulated chromium

present at time t

OL, a = the fraction of the total body content at t = 0, retained via the fast and the slow phase resp.

t = retention time t »»

Ti , Ti = the half period of the fast and the slow phase resp.

The halfperiods for both phases as well as the fractions (a , a ) appeared to be independent of the exposure concentration, length of exposure and total body content at t = 0. It was found that 34% of the total amount Cr(VI) accumulated, was retained with a halfperiod of 1.0 day and 66% with a half period of 25.6 days. Thus the actual retention equation could be obtained.

Actual retention equation:

R(t) = 0.34 . e " 0 · 6 9 3 t + 0.66 . β"0'027 t

During the excretion period the chromium content in the digestive tract decreased rapidly, while the chromium concentration in gills, liver and remainings (muscle, skin etc.) decreased less significantly.

A tentative model, describing the excretion of chromium from the tissues was designed, based on these tissue distribution data.

REFERENCES

(1) Perl, W., A method of curve fitting by exponential functions. Int. J. Appl. Rad. Isotope 8, 211-222 (1960).

The Influence of the PH on the Toxicity of Hexavalent Chromium to Rainbow Trout (Salmo Gairdnerii)

A. S. HOGENDOORN-ROOZEMOND*, V.J. H. M. TEN HOLDER*, J. J. T. W. A. STRIK*, Z. KOLAR** andj. H. KOEMAN*

^Department of Toxicology, Agricultural University, Wageningen **Interuniversity Reactor Institute, Delft, The Netherlands

The TL values of sodium chromate solutions were determined for young rainbow trout vO.10-0.30 g) at different pH levels (the water temperature was kept at 12°C and the total hardness was 1.55 meq/1). The toxicity of chromates was found to be pH dependent: a 50 to 200 fold increase in toxicity was observed when the pH decreased from 7.9 to 6.8. The toxicity data are presented in Table 1.

TABLE 1 The TLm Values of Sodium Chromate Solutions for Young

mean gH

7.9 6.8

7.9 6.8

7.9 6.9

7.9 6.9

Rainbow

gH range

7.8-8.0 6.4-6.9

7.8-8.0 6.4-6.9

7.8-8.0 6.5-7.2

7.8-8.0 6.5-7.2

Trout at Different

exposure time

48 48

72 72

96 96

120 120

j>H_

(hrs)

Levels

Itin. cone, (pgm Cr)

58.4 0.28

36.3 0.22

32.0 0.57

28.0 0.44

An explanation for this pH effect has been sought in the ionization equilibria involved; any aqueous solution of either a bichromate or chromate salt will ionize according to the following expression.

2 Croj;" + 2H+ < > 2 HCrO~ < > Cr 0^~ + H20

+ +

2H+ H +

2 H CrO HCr*0

477

478 A. S. Hogendoorn-Roozemond, et aL.

The concentrations of these Cr(VI) compounds have been calculated for the solutions causing 50% mortality at the different pH levels after 4-8, 72, 96 and 120 hours of exposure. When these concentrations are related individually to the observed mortality rates, CrO^ is found to be the most probable Cr(VI) compound responsible for hexavalent chromium toxicity to young rainbow trout.

In order to verify the often reported difference in toxicity between bichromate and chromate (Ref. 1, 2) their toxicity was compared in TI^ experiments. No difference in toxicity between these two salts was observed when the pH was kept constant at 7.9 in both solutions. However, when the pH in the bichromate solution was not adjusted and as a result decreased, mortality was found to increase. Thus, the generally found higher toxicity of bichromate may well be due to the decrease in the pH resulting from dissolving this salt in water.

REFERENCES

(1) Trama, F.B. and Benoit, R.J., Toxicity of hexavalent chromium to bluegills (Lepomis maovochirus). J. Water Pollut. Control Fed. 32, 8 (1960).

(2) Doudoroff, P. and Katz, M., A critical review of literature; toxicity of wastes to fish II. Sewage and Industrial Wastes 25, 7 (1953).

Sublethal Effects of Rhinewater on Rainbow Trout

C. L. M. POELS

Testing and Research Institute of the Netherlands Water Undertakings, KIWA Ltd., Rijswijk, The Netherlands

INTRODUCTION

The aquatic environment of the river Rhine is subject to a heavy pollutant loading. It is difficult to characterize Rhinewater with respect to its toxi-cological quality because of the vaste variety of pollutants, changing water discharges and seasonal influences. The last one hundred years dramatic changes in number of aquatic species, both vertebrates and invertebrates, occurred, leading among others, to the disappearance of most of the salmonlike fish species (Ref. 1,2). Several rea-sons can be mentioned to explain this phenomenon, being (1) the regulation and canalisation of the river,(2) increasing domestic wastewater discharges, (3) industrial pollution and (4) intensive professional fishcatchments (Ref.2). To gain more evidence that the chemical pollution, at least partly, is respon-sible for the disappearing of the salmonlike fish species, ue performed a longterm study after the effects of Rhineuater on Rainbow Trout. In fact in our investigations ue used water from the river Lek, one of the most important tributaries of the Rhine in the Netherlands. Generally the quality of the wa-ter from the river Lek is better than from the Rhine itself.

GENERAL DESCRIPTION OF THE EXPERIMENTAL DESIGN

Rainbow Trout (Salmo Gairdneri, R.) were purchased from a trout hatchery and kept in quarantaine during 2 months prior to the experiment. During the expe-riment the trout were kept in 500 1. stainless steel flowthrough basins on Rhinewater and compared with a control group on a very good quality of ground-water under identical experimental conditions. The Rhinewater was untreated except for a sedimentation period of 1 hour. The water-flow was 2 to 8 l/min and the water-temperature in the testbasins of both groups followed the river-water temperature. The water was well aerated with compressed air. The fish were fed with artificial food (Trouvit). For both groups the amount of food per kg fish was identical as well as the amount of fish per liter water. The experiment started in October 1975 and finished 18 months later in April 1977. At three months intervals samples of 20 to 30 fish of each group were investigated for eventual sublethal effects by comparing the Rhinewater group with the control group (Ref. 3).

RESULTS AND DISCUSSION

The results after 18 months exposition of the Rainbow Trout to either Rhine-or groundwater are shown in Table 1.

479

480 C. L. M. Poels

TABLE 1 Effects on a Number of Parameters in Rainbou Trout after 18 Months Exposure to Rhineuater as Compared uith Controls

Parameter

(mean of 20 to 30 fish, uith st

control

andard error of

Rhineuater

an) age of ntrol

bodyueight (g) somatic liver index (1) somatic kidney index (1) somatic spleen index (1) hemoglobin content blood (g/100 ml) hematocrit value (vol %) glucose content blood (mg/100 ml) ureum content serum (mg/100 ml) APDM activity per mg liver protein APDN capacity per liver (2) APDM capacity per liver relative

to body weight (2) GPT of liver (mU/ml) (3) LDH of serum (mU/ml) (4) AF of serum (mU/ml) (5) Acetylcholine esterase activity in

brain tissue (mU/mg protein)

( 2 )

6 3 1 , 5+. 1 ,13+0 ,23 0,75+J3,19

0 , 1 0 9 ± 0 , 0 3 9 4,32+Ό,65 4 7 , 2 + 7 , 3 41,9+_9,7

0,7+_2,1 4 , 8 1 ± 2 , 1 3 3832+J658

6 , 3 4 + 2 , 5 2 627+_218 627+_385 201+/I23

88,9^+14,8

370 ,3 1,50 0,85

0 ,148 2 ,95 35 ,9 62 ,5

7 .3 4 ,18 2688

*** +_

1+0,21 i ± 0 , 1 7 1 + 0 , 0 5 5 * *

' V 7 , 3 * * * i + 1 1 , 8 * * * '±19 8* Γ+1,64 + 1240 HK

7,16+_2,83 862+J259** 940+.501* 250£74 65,3+_10,9**

41,4 32,7 13,3 35,8 31,7 23,9 49,2 16,2 13,0 29,9

12,9 37,5 50,1 24,4 26,6

tissue weight tö (1) somatic tissue index: , . . . . body weight

(2) Aminopyrinedemethylase activity express nmoles formaldehyde produced p. hour

(3) Glutamate pyruvate transaminase (4) Lactate dehydrogenase (5) Alkaline fosfatase In addition there uas a significant accumul the most important being hexachlorobenzene, (resp. 6, 2,5 and 2 ppm in fresh addipose t groups uere present under identical experim uater, ue conclude that the Rhineiuater cont serie of sublethal effects on rainbou trout between October 1974 and Dune 1975 shoued s majority of the sublethal effects are chara experimental periods. The results furtheron tants present in Rhineuater are, at least p appearance of the salmonlike fish species f

x100

ed as

ation of pentachl issue)· B ental con ained sub An earl

imilar re cteristic suggest

artially rom the r

KK P <0,1

P <0,05

P <0,005

chlorinated hyd orobenzene and ecause both exp ditions in uell stances uhich c ier 9 months ex suits suggestin for Rhineuater

that the chemic responsible for iver Rhine.

rocarbons, PCB!s erimental aerated aused a periment g that the in the

al pollu-the dis-

REFERENCES

(1) C.3.A. van liiijck, Onderzoek naar de visfauna in de omgeving van Nijmegen Report no.32, Zool.Lab.Univ.Nijmegen. The Netherlands (1971)

(2) Deutscher Bundestag, 7. Wahlperiode.(1976) UmvUeltprbbleme des. Rheins. Bonner Univ. Buchdr., Bonn.

(3) C.L.n. Poels and 3.3.T.U.A. Strik (1975), Chronic toxic effects of the uater of the river Rhine upon rainbou trouts, in Sublethal Effects of Toxic Chemicals on Aquatic Animals (3.H. Koeman and J.D.T.lii.A. Strik ed.). Elsevier, Amsterdam, pp. 81-91.

ACKNOWLEDGEMENT

This uork is part of the research program of the KIUA and financed by the Netherlands Uateruorks Association (WEUJIN).

Studies on Aquatic Pollutants in Refotion to Neoplastic Disease of Marine Animals

ISAO TOMITA, NAOHIDE KINAE and SHINICHI SAITOU

Shizuoka College of Pharmaceutical Sciences, 2-2-1 Oshika, Shizuoka, Japan (422)

Nibea mitsukurii with neoplastic disease is often found along the coastal area of the south-west districts of Japan (Ref. 1). In a recent survey by I.Kimura, 47.4% of 1935 specimens from Kumano was found bearing melanoma.

The concentration of contaminants including chlorinated hydrocarbons (BHC,DDT and PCB), phthalic acid esters(DNBP and DEHP), inorganic and organic mercury in the edible parts of N.mitsukurii at Kumano was within the same range as found in those of other species of marine animals. The level of benzo(a)pyrene, however, was abnormally high in several cases(0.32-4.41 ppb wet weight) and this appeared to be reflected by the high level of benzo(a)pyrene in the sediment samples obtained from Kumano district(0-50.8 ppb wet weight).

By incubating with liver S-9 fraction of N.mitsukurii, benzo(a)pyrene was transformed to the substance(s) mutagenetically active to Salmonella typhimurium TA 153 8. The activity of arylhydrocarbon hyd-roxylase(AHH) analogous to that found in mammals was detected in the liver of N.mitsukurii. The enzyme activities of the fishes caught at Kumano and Kouchi, were 0-11.8 and 0-20.4(p moles of 3-hydroxy benzo(a)pyrene formed/mg prot.) respectively, while no activity was detected in N.mitsukurii trapped in other areas where very low frequencies of neoplastic dis-ease were recorded. Although AHH activity in N.mitsukurii is low compared to that in Cyprinus oarpisΛ the oral administration of PCB induced the enzyme activity(150 mg/kg body weight administration of Kanechlor 500 increased AHH activity from 0 to 60.2 p moles/mg prot.).

Another investigation was undertaken in an attempt to explore the substance(s) other than the above contaminants responsible for the neoplasm in the fish. Sediment and waste water samples collected from Kumano district were fractionated and submitted to Rec and mutagenic assays with B.subti-lis H 17(Rec+), M 45 (Rec"") or E.ooli B/r WP-2> try' hcr~ respectively. The results are shown in Table 1.

481

482 I. Tomita, N. Kinae and S. Saitou

TABLE 1 Yields and Mutagenesis of Fractions from

Fractions from sediment or waste water

neutral

acidic

basic

amphoteric

phenolic

Sediment and

Yield mg/kg

136.9

6.7

2.2

5.2

9.8

ί waste

or£

(2.1)

(5.0)

(1.1)

(0.3)

(2.8)

water

B, H . s u b t i l i s 17 M 45

* (")

± (-)

* (±)

- (i)

± (-)

E. eoli B/r WP-2 try- her"

- (-)

- (-)

+ ( + )

■H ( + )

- (")

numbers in parentheses indicate the results from waste water

Ether extract of the liver from N.mitsukurii at Kumano district, purified through silica-gel column, produced DNA damage to B.subtilis M 45. Analyses with TLC, HPLC and MS are in progress to identify the substance(s) found mutagenic in the above microbilogical assays.

We appreciate Dr.I.Kimura of Aichi Cancer center for his invaluable advice and help in this investigation. This study was supported by the Science and Technology Agency of the Japanese Government.

REFERENCE

(1) I.Kimura, Studies on tumor formation: tumores in low vertebrate animals, IGAKU NO AYUMI 96, 216 (1976).

Partially Induced Hepatic Mixed-Function Oxidase Systems in Individual Members of Certain Marine Species from Coastal Maine and Florida

JOHN R. BEND, GARY L. FOUREMAN and MARGARET O. JAMES

Laboratory of Pharmacology, National Institute of Environmental Health Sciences, Research Triangle Park, North Carolina 27709, USA

ABSTRACT

Using elevated hepatic microsomal benzo(a)pyrene (AHH) hydroxy-lase activities and/or inhibition of AHH activity by a-naphtho-flavone (α-NF) in vitro as criteria for induction, 11 of 13 winter flounder (Pseudopleuronectes americanus) and 8 of 81 sheepshead (Archosargus probatocephalus) tested had partially induced hepatic mixed-function oxidase systems. Such induction was not observed in any of the little skates (Raja erinacea) or dogfish sharks (Squalus acanthias) studied.

INTRODUCTION

The induction of hepatic AHH activity in fish has been suggested as a monitor for petroleum pollution in the marine environment (1). While this suggestion has considerable merit, some caution must be exercised in the interpretation of such data for several reasons. First, chemicals other than polycyclic hydrocarbons occur as aquatic pollutants and dramatically induce AHH activi-ties in various fish species tested. Dioxins, certain poly-chlorinated biphenyl (PCB) isomers, and certain polybrominated biphenyls (PBB) isomers fall into this category (2). Second, not all fish species respond to polycyclic hydrocarbon pretreat-ment. For example, repeated administration of 3-methylcholan-threne (3-MC) to the Atlantic stingray (Dasyatis sabina) did not affect hepatic AHH activities (3). Furthermore, if hepatotoxic contaminants are present, AHH activities may be lower in fish collected from polluted areas than in those captured from more pristine waters (4).

However, there remains little doubt that enzyme induction in aquatic species can potentially indicate the presence of selected toxic chemicals in both freshwater and marine environments. In this study, we have determined hepatic microsomal AHH activities in the presence and absence of α-NF in four marine vertebrate species from coastal Maine or Florida to test for induction.

483

484 J . R. Bend, G. L. Foureman and M. 0. James

METHODS

Adult male or female little skates, winter flounder, and dogfish sharks were captured near Mount Desert Island, Maine. Sheeps-head were collected near Marineland, Florida. Specimens were acclimated in aquaria or pools equipped with continuously circu-lating seawater or livecars immersed in saltwater for at least 24 hr before use near the site of capture. Liver microsomes were prepared, the protein content determined, and AHH activities measured exactly as described earlier (5) except that additional incubation mixtures containing 10~" M or 10~ M α-NF (in vitro) were assayed simultaneously. Incubation temperatures were 30° for the three Maine species and 35° for the sheepshead.


Pretreatment of each of the four marine species utilized in this study with polycyclic aromatic hydrocarbons (PAH; 1,2,3,4-di-benzanthracene or 3-MC) resulted in significant increases (up to 40-fold) of hepatic AHH activity. Moreover, α-NF significantly inhibited AHH in microsomes from the PAH-pretreated fish whereas it stimulated AHH activity in almost all untreated fish (2, 3). By analogy to previous studies in mammals this data is consistent with the formation of cytochrome P-448 in fish following PAH ad-ministration (6, 7). Indeed, we have recently identified cyto-chrome P-448 in hepatic microsomes of PAH-induced little skates, although it was not the predominant form of the cytochrome (8).

As shown in Table 1, hepatic microsomal AHH activity varied markedly in flounder assayed in Maine from June-August, 1977 (over 60-fold). This variation in enzyme activity is likely an indication of some induced AHH activity in certain fish, although it could also be related to the heterogeneity of the wild fish population. In only 2 of the fish studied did α-NF stimulate hepatic microsomal AHH activity whereas inhibition was observed in all other fish. Since the activitation caused by a-NF occurred in the 2 fish with lowest hepatic microsomal AHH acti-vities, this also suggests that the majority of the flounder tested had partially induced mixed-function oxidase systems. It is attractive to assume that this is due to exposure to environ-mental pollutants, although this is not definitely known to be the case. In contrast, the hepatic AHH activity of the 7 dogfish sharks and the 7 little skates tested was stimulated by in. vitro α-NF (data not shown), consistent with no induction of the hepatic AHH system.

Of 81 sheepshead studied in Florida only 8 fish had hepatic microsomal AHH activity that was inhibited by α-NF, and most of the inhibited fish (85%) were assayed during April and May, which coincides with the spawning season. This suggests that normal physiological processes may also be related to the partially in-duced AHH activities observed in these sheepshead.

Partially Induced Hepatic Mixed-Function Oxidase Systems 485

TABLE 1 Hepatic Microsomal Benzo(a)pyrene Hydroxylase Activity, in Presence and Absence of 10" M α-Naphthoflavone, in Winter Flounder (P. ameri-canus) Collected from Coastal Maine

AHH ACTIVITY

Fish No.

1 2

3 4 5 6 7 8 9

10 11 12 13

Without a-NF

0.05 0.26

0.28 0.40 0.82 1.75 1.76 1.93 2.07 2.09 2.80 2.99 3.15

With a-NF

0.19 1.12·

0.21 0.19 0.45 0.61 0.83 1.19 1.63 1.38 1.63 1.53 0.80

Fluorescent units/min/mg microsomal protein.

CONCLUSIONS

Collectively, the data demonstrate that some individual flounder and sheepshead captured in the wild have partially induced mic-rosomal mixed-function oxidase systems. However, whether or not the induction in these two species is due to exposure to cer-tain classes of environmental pollutants or is related to normal physiological changes is yet to be definitely elucidated.

REFERENCES

(1) J. F. Payne, Field evaluation of benzpyrene hydroxylase in-duction as a monitor for marine petroleum pollution, Science 191, 945 (1976).

(2) J. R. Bend and M. 0. James, Xenobiotic metabolism in fresh-water and marine species. In: D. C. Malins and J. R. Sargent (eds.), Biochemical and Biophysical Perspectives in Marine Biology, Academic Press, New York. In. press.

(3) J. R. Bend, M. 0. James, and P. M. Dansette, In vitro metabolism of xenobiotics in some marine animals, Ann. N. Y. Acad. Sei. 298, 505 (1977).

(4) J. T. Ahokas, N. T. Karki, A. Oikari and A. Soivio, Mixed-function monooxygenase of fish as an indicator of pol-lution of aquatic environment by industrial effluent, Bull. Environ. Cont. Toxicol. 16, 270 (1976).

(5) R. J. Pohl, J. R. Bend, A. M. Guarino and J. R. Fouts, Hepatic microsomal mixed-function oxidase activity of several marine species from coastal Maine, Drug Metab. Disp. 2, 545 (1974).

486 J. R. Bend, G. L. Foureman and M. 0. James

(6) F. J. Wiebel, J. C. Leutz, L. Diamond and H. V. Gelboin, Aryl hydrocarbon (benzo[a]pyrene) hydroxylase in micro-somes from rat tissues: Differential inhibition and stimulation by benzoflavones and organic solvents. Arch. Biochem. Biophys. 144, 78 (1971).

(7) F. J. Wiebel and H. V. Gelboin, Aryl hydrocarbon (benzo[a]-pyrene) hydroxylase in liver from rats of different age, sex and nutritional status: distinction of two types by 7,8-benzoflavone, Biochem. Pharmacol. 24, 1511 (1975).

(8) T. H. Elmamlouk, R. M. Philpot and J. R. Bend, Separation of two forms of cytochrome P-450 from hepatic microsomes of 1,2,3,4-dibenzanthracene (DBA) pretreated little skates (Raja erinacea), Pharmacologist 19, 160 (1977).

Organics in Air, Rain, Snow and Lake Surface Water

B. VERSINO, H. KNOEPPEL and H. VISSERS

Commission of the European Communities, Joint Research Centre — ISPRA (Varese) Italy

During our experiments dealing with particle formation in air, we tried to correlate the organics present in other media such as rain, snow, lake surface water· During the last winter we started some preliminary experiments which are reported here.

Samples have been collected in a semi-rural area, which, especially during the winter time, is mostly polluted by gasoline car exhaust gases. Sampling of air and water as well as analysis by GC-MS-computer system and identification by mass spectrum library search techni-ques have been described elsewhere (1, 2).

Fig. 1 shows the flame ionization detector capillary gas chroma-togram of an air sample (40 liters) taken during a snowfall, of the fresh snow itself (1 1 of melted snow), of rain (1 1, mixed with some snow) sampled some days after the snow fall and of the surface water of a nearby lake. The air sample contains, as usually, predominantly organics originating from car exhaust with toluene (peak B ) as the most important constituent. (Just for sake of comparison, peak A corresponds to benzene, peak C to tetrachloroethylene, the group of peaks D to ethylbenzene and the xylene isomers.) Snow contains essentially the same compounds like air but, due to an obvious washout effect, higher boiling compounds are more pronounced. This is even more evident from the snow plus rain sample. For a comparison of the air, the snow and rain chromato-grams it should, however, be noted that approximately 50 g of air but 1000 g of snow resp. rain have been sampled. The lake surface water chromatogram is more similar to the air one and contains, for instance, also a very pronounced toluene peak. The higher boiling compounds present in snow and rain are also present but are obviously much more diluted.

487

488 B. Versino, H. Knoeppel and H. Vissers

Fig. 2shows the electron capture detector capillary gas chroma-tograms of some of the samples in the previous figure. The air contains mostly the chlorinated compounds such as chloroform, 1,1,1-trichloroethane, carbontetrachloride, trichloroethylene, tetrachloroethylene (respectively peaks 1, 2, 3, 4, 5) and practically no electron capturing compounds in the higher boil-ing, range (this part of the chromatogram is omitted). The snow and rain sample contains the same chlorinated compounds present in the air plus many others in the higher boiling range. Again, as for the flame ionization trace, the lake surface water looks more "clean" than the snow, although to a lesser extent in the case of the electron capturing compounds.

Fig. 3 shows the flame ionization and electron capture detector traces of another snow sample. Again it is interesting to note that all the compounds belonging to car exhaust (for instance peaks A, B, C, D) and all the principal chlorinated compounds such as chloroform, 1,1,1-trichloroethane, carbontetrachloride, trichloroethylene, tetrachloroethlene (respectively peaks 1,2, 3,4,5) are present and that the snow contains many other electron capturing compounds the structure of some of them we intend to clarify during the next winter.

Anyhow, even from this preliminary results, it is clearly indi-cated that, by snow and rain, air pollutants are washed out, with an enrichment of the higher boiling compounds, are trans-portated to the ground and, hence, are likely to enter and contaminate water sources. Compounds such as benzene, trichloro-ethylene, tetrachloroethylene, which are to-day of great concern for their carcinogenity, can enter the water not only by spillage but, particularly in rural and semi-rural areas, by transport and wash-out from air.

This indicates that, food apart, water and air standards should be defined jointly taking into account the accumulated man daily intake of hazardous organic compounds and, hence, the associated risk.

References

(1) B. Versino, H. Knoeppel, M. De Groot, A. Peil, J. Poelman, H. Schauenburg, H. Vissers and F. Geiss J . Chromatography 122 ( 1 9 7 6 ) , 373-388

(2) H. Knoeppel, B. Versino, W.G. Town, H. Schauenburg, A. Peil, J. Poelman, F. Geiss and 0. N^rager Advances in Mass Spectrometry, Vol. 7 (1977), in press.

Organics in Air, Rain, Snow and Lake Surface Water 489

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Organics in Air, Rain, Snow and Lake Surface Water 491

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The Significance of Surface Microlayers in Evaluation of Persistent Pollutants with Aquatic Laboratory Model Ecosystems

ANDERS SÖDERGREN

Institute of Limnology, S-220 03 Lund, Sweden

It is important that as much information as possible is gathered when

the pattern Of distribution and mechanism of degradation of persistent

compounds are studied in laboratory model ecosystems. This usually im-

plies sampling and analysing various biotic and abiotic components

within the system for residues of the compound tested. However, the

surface microlayer is one component which is seldom recognized and in-

cluded in the ordinary sampling procedure. Therefore, I would like to

stress the importance of the surface microlayer in aquatic systems in

general and its significance in model ecosystems in particular.

Freshwater and marine surface films have physical, chemical and bio-

logical properties quite different from those of the water underneath.

Baier et al (1974) argue that lipids are rare in surface films except in regions of pollution, while other workers (Garrett 1976, Larsson

et al 1974) report fatty acids and other lipids in relative large

amounts. Part of this controversy reflects difficulties in sampling

the surface microlayer, and it is obvious that a technique does not

exist today that will pick up all the substances which collect in the

microlayer.

In natural freshwater bodies the surface film of the interface is nor-

mally monomolecular. In lakes with high production or in polluted

areas the film may consist of several layers of hydrophobic molecules.

Due to the lipophilic character of the microlayer, it accumulates var-

ious pollutants and may also affect their degradation and transport to

and from the water body. Therefore, vital information may be retrieved

from the surface microlayer.

493

494 A. Sodergren

The green alga Cklorella pyrenoidosa was grown in continuous flow cul-

tures and fed to compartments with secondary consumers (Daphnia and

fish). The distribution of hexachlorobensene, hexachlorobiphenyl, Clo-

phen A 50 and DDT added to the system was studied. The surface film

was collected with a teflon plate (Larsson et al 1974) and with a

technique that allowed no algae to interfere with the sampling proce-

dure (Södergren, in press). Transport from the film by jet drops crea-

ted by bursting bubbles in the microlayer was followed by collecting

the drops on teflon plates 5 cm above the surface of the water. Sub-

stances on the teflon plates were extracted with hexane and the ex-

tract examined by autoradiography on TLC plates and by gas chromato-

graphy.

The major part of the substances added to the system was taken up by

the algae. However, relatively large amounts were found in the surface

film (0.5% - 3% of the amount added). Except for DDT, none of the sub-

stances tested showed a different pattern of degradation in the sur-

face film compared to other components of the system. In the case of

DDT, proportionally more DDD was found in the surface film than in

other components. This may be an effect of the bacterial population

of the interface.

All of the substances studied were transported at various rates to the

air within the system, mainly by the jet drop mechanism. Depending on

the size of the air bubbles produced and the air flow rate a maximum

of 7% (Clophen A 50) of the amount added left the system via the jet

drops. It is obvious that this mechanism of transport has to be con-

sidered when evaluating the distribution of persistent compounds in

aquatic environments. REFERENCES

Baier, R.E., Goupil, D.W., Perlmutter, S. and King, R. Dominant

chemical composition of sea-surface films, natural slicks and foams,

J.Rech. Atmosph. 8, 571 (1974).

Garrett, W.D. The organic chemical composition of the ocean surface,

Deep-Sea Research 14, 221 (1976).

Larsson, K., Odham, G., and Södergren, A. On lipid surface films on

the sea. I. A simple method for sampling and studies of composition.

Marine Chem. 2, 49 (1974).

Södergren, A. Composition and properties of surface films produced by

Chlorella pyrenoidosa, Proc. SIL, Sandefjord 1976 (in press).

Reactions of Ozone with Organic Micro-Pollutants During the Preparation of Drinking Water

E. DE GREEF

National Institute for Water Supply, 2260 AD Leidschendam, The Netherlands

In an orientating chemical-analytical survey the secondary aspects of the ozonisation of drinking water were studied. In order to investigate the influence of the ozoni-sation process on organic products formed during the chlorination of drinking water two purification plants were selected (fig. 1). which were comparable with respect to the source of the raw water (River Rhine).

One of the plants treats the water, after bank infiltration, with simple filtration and ozonisation. The purification scheme of the other plant consists of flocculation-filtration, breakpoint chlorination and ozonisation.

Water samples were taken before and after the ozone treatment and analysed for the nature and concentration of the organic compounds. The extraction, concen-tration and analysis were executed in the Organic Analytical Laboratory of the In-stitute by means of the following techniques:

CONCENTRATION + EXTRACTION - liquid/liquid extraction: pentane-ether - headspace analysis : 100 ul vapor injection - XAD-concentration : ether eluate, concentration by evaporation (200 : 1) - Grob extraction : concentration by vapour recycling CS2 eluate

QUALITATIVE + QUANTITATIVE ANALYSIS - gaschromatography : Carlo Erba, 50 m glass capillair (OVl, UCON) - mass spectroscopy : Finnigan Quadrupole, WDV Data-system

In fig. 2 the results of quantitative analysis have been compiled for some groups of organics.

495

496 E. de Greef

From this the following can be inferred: 1. Generally the ozonolysis of organic micropollutants seems to follow the reaction

mechanisms known from laboratory experiments. This means that the rules of selec-tion valid for substituents are maintained. This results in a very slow destruc-tion of multi-halogenated aliphates and aromates. The haloforms caused by the breakpoint chlorination are not removed by ozone.

2. When the ozonisation is preceeded by a breakpoint chlorination a rise in the concentration of a number of organo-halides can be ascertained, generally multi-chlorine and -bromine methyl and ethyl compounds. Presumably this is caused by ozonolysis of material with a higher molecular weight, which is chlorinated and/ or brominated during the breakpoint chlorination.

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Organic Compounds of a Possible Biological Nature Present in Tapwater of The Netherlands

G. J. PIET, C. F. H. MORRA, A. DEN BOER and B. C.J. ZOETEMAN


Organic chemicals in tapwater in The Netherlands originate from industrial sources or from the water treatment process, but other organic constituents from partly unkown sources are present too. They can be formed during riverwater storage in open reservoirs, or during the stay of water in the ground, including dune infil-tration and bankinfiltration of riverwater.

Groundwater and bankinfiltrated water Water abstracted from the soil may contain a series of relatively simple oxygen containing branched chain compounds such as, methylbutanoate, methylisobutanoate and a-methyl - methylbutanoate which occur in concentrations ranging from 0,1-1 ug/1. Other oxygen containing substances present in groundwater are: butylformate, ethy1-sec-butylether, di-n-butylether, 1,1-dimethoxypropanal, 1,1-diethoxyethane, 1,1-dimethoxyisobutane, 2,3-dimethyl-2,3-butanediol.

Several compounds present in groundwater could not clearly be identified but showed resemblance to lower ethers and esters. These substances are of interest because they might contribute to a good water quality as most of them have an agreable odour and some are supposed to have a rather low odour threshold concen-tration .

In groundwater several ketones such as heptanon-3 and nonanon-5 and several aldehydes such as butanal, 2-ethylhexanal, benzaldehyde, dimethylbenzaldehyde and cinnamaldehyde (C9H8O) occur too. Saturated and unsaturated alcohols (Cg-Cg) were also identified. Also some of these compounds have very low odour threshold concentrations.

An extensive research of a mixture of bankinfiltrated Rhine water and ground water showed that a series of mono-, di-, tri- and tetraethylene glycol, di-methyl and di-ethyl ethers were present in tapwater derived from those sources. In bankinfiltrated Rhine water several oxygen containing compounds showed great similarity to compounds such as 2-ethyl-4-methyl-l,3dioxolane (C5H12O2)r 2-methoxy-ethylacetate (C5H10O3) , 2-methyl-pentanol-2 (CgHÔ) , dimethyldioxane (C6Hi2°i / 1,3-diethyl-l,4-epoxybutane (CgHiÔ), 2-ethylhexanal (CgHÔ) and cyclodecanol (C10H20O).

Reservoir water Another group of compounds present in tapwater is possibly introduced into the water during storage in open reservoirs. In a 20 city survey in The Netherlands (1976) tapwater from different raw water sources and different water treatment systems was analysed. Several compounds seemed to be present mainly in cases where

499

500 G. J. Piet, C. F. H. Morra, A. den Boer B. C. J. Zoeteman

the raw water is stored in open reservoirs. Several of these substances probably contribute to impaired water taste (see table 1).

TABLE 1 Organic compounds, possibly introduced into the water during storage in open reservoirs

Compound Max. Cone, ug/liter O.T.C. ug/liter

1-Octene 1-Nonene 1-Undecene C8 and Cg saturated and unsaturated alcohols 2-Methylisoborneol Geosmin Hexanal Heptanal Octanal Nonanal Decanal Undecanal 2-Dodecene-4-one

0,03 0,01 0,01 0,01-1,0 0,03 0,03 0,03 0,1 0,03 0,1 0,1 0,03 0,03

0,5 ---

0,02 0,015 4,5 3,0 0,7 1,0 0,3 5,0 -

The analytical system used, which was equipped with high-resolution glass capillary gaschromatography columns which were directly connected to a fast scanning quadru-pole mass-spectrometer (2-3 scans/sec), was not able to identify 30-40% of the eluting compounds so more organic substances which have not been identified might contribute to impaired water quality. Besides this, odour threshold concentrations of many organic compounds are still unkown and sufficient knowledge concerning the mechanism and possible synergism of complex mixtures of odorous organic constituents of drinking water is still lacking.

Biological Monitoring Based on Fish Respiration for Continuous Water Quality Control

W. SLOOFF


INTRODUCTION

Since the demand for water for domestic and industrial purposes exceeds the supply of available high quality sources, greater use is being made of polluted surface water as a raw water source for drinking water. The increased pollution of these waters has led to the development of more advanced methods for pollution detection aimed at the protection of human health. Most pollution monitoring programmes are still orientated solely towards chemical and physical methods. However, these methods do not meet the requirements for an operational surveillance system related to toxic compounds (Poels, 1975) . Since living organisms will respond to every pos-sible substance or mixture of substances at some level, no matter what their chemi-cal or physical characteristics may be, the biological techniques are playing an increasingly important role in predicting and controlling water pollution (Cairns, Dickson and Westlake, 1977). Recently, several biological monitoring systems have been reviewed (Slooff, in print). From a practical point of view and re-lated to the protection of the human health some of these systems show a promising future, particularly the systems based on the relationship between water pollution and respiratory activity in fish. Although several automatic recording devices to detect fish respiration patterns have been prepared (Heath, 1972), the method using dual external electrodes is considered to have the best application possibi-lities since the fish is allowed a maximum amount of freedom under test conditions (Spoor, Neiheisel en Drummond, 1971; Waller and Cairns, 1972; Morgan, 1976). The purpose of this paper is to contribute to a final operational stage of this par-ticular system and to attempt to delineate the limit of respiratory response to various concentrations of several toxicants under laboratory conditions.

MATERIALS AND METHODS

Testconditions As biological indicator rainbow trout (Salmo gardnerii) was used, obtained from a commercial fish hatchery. The mean weight of the fish used in the experiments was 56 g (s=26). After being transported in oxygenated plastic containers to a fish stock holding facility the fish were maintained under laboratory conditions for at least two weeks prior to being used in the experiments. To obtain testwater, dechlo-rinated tapwater of constant temperature was diverted into a storage basin in which the water was aerated by chemically washed ambient air. From this basin the water was pumped by a glaspump into a circulation system in the laboratory compartments where the water was used for the fish holding facilities and experiments.

501

502 W. Slooff

The characteristics of the water were as follows:

temperature : 20 +. 1°C pH : 8,0 +0,2 dissolved oxygen content : 9,0 +_ 0,5 mg/1 hardness :10,1 +0,1 D°

All experiments were carried out under conditions of a fixed natural photoperiod to allow diurnal variation in respiration patterns. Dawn occured gradually between 6.00 and 7.00 hour, dusk between 20.00 and 21.00 hour. Each fish was kept in a testcham-ber of approx. 2,5 1. A flow rate of 30 1. hr"1 was maintained thoughout.

To minimize disturbances to the fish a separate experimental room was construc-ted which was closed during the experiments. The waterflow and the toxicant addition were regulated outside this room. To avoid responses due to floor vibrations the test chambers were placed on a heavy iron table which rested on a separate concrete foundation. Further all electrical and water connections were flexible to minimize the transmission of wall vibrations. During the experiments the fishes were not fed.

The monitoring apparatus The design of the experimental chambers was based on that of Spoor et. al. (1971), consisting of a glass tube in which two stainless steel mesh electrodes are placed at the opposite ends, covering the total cross-section of the chamber. Of each of these mesh electrodes, which detect the breathing signals, one wire is connected to a shielding cable leading to a high quality frequency selective pre-amplifier of variable gain (model 133, Princeton Applied Research Corporation). This amplifier is of the differential type to minimize the hum and noise pick-up. The amplification of this instrument is selectable between 10 and 10.000, in a 1-2-5 sequence, and an uncalibrated gain vernier provides 1 to 2.5 range expansion. The amplifiers were operated at a gain of approx. 10.000 amplifying frequency signals (in the order of 20-100 uV) between 0.3 and 10 Hz.

Subsequently the amplified signals of each fish alternately pass, via a multi-plex time switch system, to a frequency filter.,This filter consists of variable solid state electronic filters with digitally tuned cut-off frequencies over the range from 0.001 Hz to 99.9 kHz (model 3342, Kröhnhite Company). To suppress pos-sible disturbances or large variations in the level of the recorded baseline caused by body movements of the fish, the filter was operated with low and high cut-off frequency limits from 0,5 Hz to 5 Hz.

During the experiments, the data were collected by a multichannel paperrecorder (Minograf-800, Elema Schönander); at the present stage the system is automatized using a timer counter (5308A, Hewlett-Packard) and a calculator (9830A, Hewlett-Packard) . For each fish the respiration frequency was counted manually during one minute every hour. Since the respiration patterns at different times of the day for the same fish, and at the same time of the day for different fish are heteroge-neous, each fish had to be used as its own standard to allow for individual varia-tion. To compensate for the diurnal variations the critical values were determined as the minimum and maximum respiration frequency during a light and dark period recorded after acclimatization of the fish to the experimental conditions. It was considered that the fish recovered from the initial stress of being netted and transferred into the test chambers within 3-4 days. After recording the standard breathing patterns, the toxicant was added continuously to the waterflow by an in-jector system (perfurer V, Braun) during 48 hours. If the respiration frequency of at least three fourth of the testfish exceeded the predetermined individual critical values at the same hourly interval, it was considered as an indication of a toxic condition.

Of each toxicant several concentrations were tested to establish the detection limit of the biological monitoring system, which was defined as the lowest concen-tration at which a toxic condition was developed within 24 hrs. after toxicant ad-ministration. To compare the obtained detection limits with common LC50 values, additional experiments with zebrafish (Brachydanio rerio) were carried out. For

Biological Monitoring 503

this purpose 10 fish were exposed to each toxicant concentration for 48 hrs. in 10 1 aquaria; the flow rate of these closed dynamic systems amounted to 6 1 hr .


In table 1 the detection limits of the biological monitoring system of 12 compounds are compared with corresponding LC50 values for fish. Apart from the fact that the LC50-values are based on an experimental exposure time twice as long as used in the biological monitoring studies, it is obvious that the described system is a much more sensitive tool in predicting developing toxic conditions of the water, than the common LC50-measurements. Morgan (1976) suggested that the response limit of fish sensors lies between 5 and 10 per cent of the 48 hour LC50, using a biological moni-tor similar to that described in this paper. The results of this study indicates a broader range of these ratio's, varying from 1 per cent (cadniium) to 30 per cent (acrylonitril, cyanide, lindane) of the 48-hour LC50. Although the difference in susceptibility to toxicants between rainbow trout and zebrafish may contribute to this phenomenon to some extent, there is no doubt that the detection capacity of the monitor system depends largely on the mode of action of the toxicants. In gene-ral fast responses of low concentrations will be observed for those pollutants which act more or less directly on the respiratory system. Other chemicals, e.g. carcinogenic compounds or other compounds which require bioactivation into reactive metabolites before exerting a toxic action, may pass unnoticed in hazardous quanti-ties (Koeman et. al., 1977). It was noticed that more responses were obtained during the dark interval than during the light interval, which is in accordance with the findings of Cairns et. al. (1973) . This may be due to the fact that the breathing rates were generally lower and often less variable during the dark interval than during the light interval resulting in close upper and lower critical values at night. Besides it is also possible that the sensitivity of the fish to the toxicants is subjected to a circadian periodicity (Cairns et. al., 1973).

In most cases the toxicants caused an increase of the breathing rate of the fish in sub-lethal concentrations, even up to 2-3 fold the normal value. Only fish ex-posed to low concentrations of acrylonitril showed first a slight but significant decrease in respiration frequency, followed by an increase later on, while exposure to chloroform showed individual differences in respiratory behaviour among the fish.

Since the biological monitoring system is aimed at the protection of the human health, the results are also compared with literature data on LD50 values for rats (Table 1). In comparison with these data, the monitor system with rainbow trout was capable to detect the selected compounds far below the critical concentration for rats. Based on a body weight of 100 g of the rat, drinking a single dose of 10 ml a theoretical mean safety factor of almost 19.000 was found, whereas only for acryl-onitril and cyanide the safety factor was less than 1000. Taken into account these results, and the fact that a comparison of the degrees of sensitivity of fish and human beings showed that fish proved more sensitive to acute intoxication by pollutants than men (Jung(1973) ; Dauson et. al., (1975), the monitor is not only suitable to prevent ecological damage by monitoring industrial and domestic effluents, but also to protect the human health by the capacity to detect harmful concentrations of many toxicants which may occur in water to be used for drinking water supply (scheme 2).

CONCLUSIONS

It can be concluded that this system might be helpful to control water quality in effluents and surface waters in addition to chemical/physical monitors. Further field studies will ultimately have to show which modifications are necessary to fit the requirements of an operational automatic biological monitoring system under local conditions. However, it should be emphasized that still hazardous quantities of certain compounds may pass unnoticed. Therefore further development of specific biological monitoring tests is needed.

504 W. Slooff

ACKNOWLEDGEMENTS

This publication is partly based on the work which was supported by the Commission of the European Communities under contract nr. 111-75-1 ENV. The author is greatly indebted to Mr. E. Rab and Mr. T.P. Spierenburg for their technical advice and to Mr. R. Baerselman for his continuous assistance.

REFERENCES

1. Cairns, J.Jr., Dickson, K.L. and Westlake, G.F. (ed)(1977) Biological monitoring of water and effluent quality ASTM Techn. publ. 04-607000-16, Symposium "Biological monitoring of water eco-systems", Blacksburg, Virginia (1975)

2. Cairns, J.Jr., Hall, J.W., Morgan, E.L., Sparks, R.E., Waller, W.T. and Westlake G.F., The development of an automated biological monitoring system for water quality Bulletin 59, Virginia Water Resources Research Center, Virginia Polytechnic Institute and State University, Blacksburg, Virginia

3. Dauson, G.W., Stradley, M.W. and Shuckrow, A.J. (1975) Determination of harmful quantities and rates of penalty for hazardous substances Vol II technical documentation EPA-440/9-75-005-6

4. Heath, A.G. (1972) A critical comparison of methods for measuring fish respiratory movements Water Research 6, 1-7

5. Jung, K.D. (1973) Extrem fish toxische Substanzen und ihre Bedeutung für ein Fischtest-Warnsystem GWF-Wasser-Abwasser, 114, 232-235

6. Koeman, J.H., Poels, C.L.M. and Slooff, W (1977) Continuous biomonitoring systems for detection of toxic levels of water pollutants Paper presented on the 2nd International Symposium on Aquatic Pollutants, September 26, 27 and 28, 1977 Noordwijkerhout, The Netherlands

7. Morgan, W.S.G. (1976) Fishing for toxicity: biological automonitor for continuous water quality control Effl. Wat. Treatm. J. 16, 9, 471-472, 474-475.

8. Poels, C.L.M. (1975) Continuous automatic monitoring of surface water with fish Water Treatm. and Exam. 24 (1) 46-56

9. Slooff, W., (1977) Een evaluatie van biologische testsystemen ter bewaking van de waterkwaliteit Will be published in H^O

10. Spoor, W.A., Neiheisel, T.W. and Drummond, R.A. (1971) An electrode chamber for recording respiratory and other movements of free swimming animals Trans. Am. Fish. Soc. 100, 1, 22-28

11. Waller, W.T. and Cairns, J.Jr. (1972) The use of fish movement patterns to monitor zinc in water Water Research, 6, 257-269

Biological Monitoring 505

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Table 1: The relative sensitivity of a biological monitoring system with dual

external electrodes etc.

Final Considerations

P. SANTEMA Director of the National Institute for Water Supply, Voorburg, The Netherlands

As a large variety of problems have been discussed during the past three days it is nearly impossible to summarize them all at this moment in an adequate way. Nevertheless I would like to mention several items, which seemed to me to be important, without pretending to be exhaustive.

In order to recognize priorities for toxicological research and sanitation matters it has been recommended that the establishment of a database on the occur-rence of chemicals in water should be strived after in a joint international effort. Different speakers noted the presence of hazardous chemicals in groundwater in The Netherlands as well as in Switzerland stressing the need to watch carefully our groundwater resources and promote better protection measures. Very interesting figures for the situation in The Netherlands showed that the dis-charge by major industries in The Netherlands will be reduced in the period 1975-1980 by 95% for mercury, by 80% for arsenic, while cadmium will not be significantly reduced. The total contribution by Dutch industries to contamination of the river Rhine by heavy metals is generally less than 10-20% of the amount imported by the rivers Rhine and Meuse.

Dr. van Lookeren Campagne, representing the Rotterdam Industries, came with a result-oriented conception of cleaning the surface waters and the Rhine in par-ticular; part of the conception brings a registration and publication of all effluents in the whole Rhine-basin. Factors influencing chemicals in the environment were discussed. Volatilisation, chemical and photochemical degradation and microbial transformations are important. Compounds belonging to the black and grey list of the Chemical Convention against pollution of the river Rhine were discussed including benzopyrene, nitrosamines and taste and odour impairing compounds.

The side effects of water treatment have been discussed. I think that all participants agree that the level of carcinogens in drinking water should be as low as possible. This means for the application of chlorination that breakpoint chlorination should be avoided as soon as possible and that application of chlori-nation for disinfection should only take place after rigorous removal of organic carbon levels. Furthermore alternative oxidants like ozone should only be considered for application after a careful screening of possible adverse side effects.

Finally it is my pleasure to announce that the 3rd International Symposium on Aquatic Pollutants will be held in the U.S.A. in the fall of 1979 under the respon-sability of the E.P.A.

507

Index

AAF 4 ABS surfactants 98 Accumulation 310-311, 399 Acetanilide 300 Acetone 247 Acetophenone 230, 232 Acrylamines 126 Acrylonitrile 343, 425, 434, 503 Adsorption 20, 175, 468

of carbon 363 Aerosols 128 Aflatoxin 3, 4, 34, 353 Air-water interface 175, 177 Alcaligenes eutrophus 193 Alcohols 499, 500 Aldehydes 98, 103, 499 Aldrin 207, 251, 427, 444 Algae 251, 494

blue-green 71, 77 green 494 red 352

Algal contribution 69 Aliphatic hydrocarbons 103 Aliphatics, chlorinated 96 Alkanes 69-82

cyclic 72, 77, 82 leaf wax 71

Alkoxy radicals 224, 230 Alkylated aromatic hydrocarbons 110,

120 Alkylbenzenes 103, 110, 114, 119, 120 Alkylbiphenyls, chlorinated 28, 212 Allochthonous contribution 70, 72, 78 Amberlite XAD-2 156 Ames test 352, 405, 432, 433, 481 Amines 97

aromatic 214, 272, 343 4-aminobiphenyl 425, 434 5-amino-4-chloro-2-phenyl-3(2H)-

pyridazinone 193 4-aminostilbene 425, 434

Ammonia 280, 343, 344, 346 Ammoniacal-N 279, 280 Ammonium concentrations 471-472 Analysis 87-89

discriminant 23 multidetection 87 of organic pollutants 39, 40, 55,

63-67 of tap water 103

Anilines 24, 126, 246, 247, 272, 273, 287, 300, 303

chlorinated 366, 367 Anionic detergents 278-279 Anthracene 192, 247 Anthranilic acid 196 Appleton, H. 473 Archosargus probatocephalus 483 Arene oxides 301, 303 Arenicola cristata 312 Aromatic amines 214, 272, 343 Aromatic hydrocarbons 103, 187, 188,

192, 193 alkylated 110, 120 chlorinated 300-302 polycyclic 126, 128-129, 132

Aromatic pollutants 187-201 chlorinated 97, 287 degradation 197 metabolism 187, 188-190

Arsenic 127, 132, 133, 136, 139, 140, 421, 422, 507

Arthrobacter 190 Aryl hydrocarbon hydroxylase (AHH)

299, 302-303, 481, 483, 484, 485 Asbestos 420 Assimilative capacity 172 Aucilla River 244, 246, 249 Autochthonous contribution 70, 71, 77 Azobenzene 427, 436 Azobis 228, 232

509

510 Index

Balb/3T3 cells 408, 409, 413 Banerjee, S. 473 Bankfiltration 265-273, 362, 499 Barium 422 Basalin 215, 216 Bathing water 331 Baughman, George L. 237 Behavioural responses 311 Beijerinckia 192 Benchmark method 17, 20 Bend, John R. 483 Benthic system 312 Benzenedihydrodiol 192 Benzenes 25, 97, 110, 178, 192, 193,

196, 247, 287, 365, 425, 443, 444, 488

alkylated 103, 110, 114, 119 chlorinated 114, 302, 367

Benzidine 425, 434 Benzo(a)anthracene 193, 194, 242,

243, 247, 425 Benzoate 199 Benzo(b)fluoranthene 425 Benzo(j)fluoranthene 427 Benzo(k)fluoranthene 427 Benzoic acid 24, 193, 196, 200

chlorinated 187, 199-201 Benzo(g,h,i)perylene 427, 437 Benzophenone 246, 247 Benzopyrene 126, 127, 128, 507 3,4-benzopyrene 126 Benzo(a)pyrene 4, 193, 194, 241, 242,

243, 247, 250, 252, 253, 300, 302, 303, 351, 411, 413, 415, 425, 434, 444, 481, 484, 485

Benzopyrene hydroxylase 353 Benzo(f)quinoline 242 Berne Agreement (1963) 326 Beryllium 132, 421, 422 BHC 427, 444, 481 Bioaccumulation 2, 6, 7, 8, 13, 14,

16, 20, 21, 23, 26, 283, 285-290, 311

Bioassays 309, 311, 369-380 models 341

Bioconcentration of pollutants 352 Biodegradation 19, 71, 74, 75, 150,

152, 283, 285, 287, 317, 398, 465

of oil 70 Biohydroxylation 300 Biological oxygen demand (BOD) 276-

280, 285 Biomonitoring systems 341-347, 501-

503 Biorefractories 419, 423, 448, 449 Biphenyls 28, 178, 192, 193, 194, 198,

247, 300 brominated 210, 212

Biphenyls (cont'd) chlorinated 197-199, 210

bis(2-chloroethyl)ether 427, 436, 444 bis(chloromethyl)ether 425, 435 Bitumen 128 Blacklist of pollutants 7, 8, 17, 153,

327, 328, 331, 507 Borneff, J. 125 Brachydanio rerio 346, 353, 502, 503 Brasseil, S.C. 69 Breakpoint chlorination 364, 365, 495,

496 Brominated compounds 421 Bromine 25 Bromobenzenes 25 Bromobiphenyls 25, 210, 212 4-bromobiphenyl 303, 304 2-bromoethylpropane 427 Bromoform 438, 446 p-bromotoluene 193 Bulk air phase processes 179-180 Bulk water phase processes 177, 179 Butylbromide 427, 436

Cadmium 136, 137, 139, 140, 422, 423, 503, 507

Cancer 391, 396, 447 lifetime risk factors 444-445 mortality rates 421

Carbamates 217, 270, 345, 346 Carbaryl 242, 243, 246, 247, 440 Carbazole 242 Carbon, organic 275-280, 467, 468

adsorption 363 budget 463 concentration 464 measurements 66, 463, 464

Carbon filtration, activated 268, 271, 363, 392

Carbon tetrachloride 4, 101, 102, 103, 108, 110, 111, 112, 113, 114, 116, 118, 120, 121, 130, 268, 349, 425, 444, 488

Carcinogenesis 303 testing 405

Carcinogenic potential 349 Carcinogenicity 2, 3, 8, 17, 18, 108,

125-127, 303, 433 testing 370

Carcinogens 3, 4, 125-133, 303, 343, 349-354, 396, 397, 420, 421, 423, 424, 431, 434-435, 441, 443, 448, 450

classification 419, 424, 431 indicators for 349 organic 423 testing for 389-390

Carp 351 Catastomus commersoni 351

Index 511

Catechol 193, 194, 196, 199 Central Water Planning Unit (UK) 384,

385, 390, 392 Chemical Substances Control Law (Japan)

29, 283-284 Chlordane 427, 436, 444 Chlorella 251 Chlorella pyrenoidosa 494 Chlorinated compounds 488 Chlorination 130, 132, 150, 349, 361,

362, 364, 365, 396, 421, 447, 495, 496, 507

Chlorine 25, 130, 137, 210 Chloroalkylenes 212 Chloroaniline 246, 273 2-chloroaniline 366 4-chloroaniline 272

Chlorobenzene 102, 103, 110, 112, 113, 114, 116, 118, 121, 150, 178, 193

metabolism 302 3-chlorobenzoate 199 3-chlorobenzoic acid 193, 199, 200 4-chlorobenzoic acid 193 Chlorobiphenyls 25, 26, 27, 28, 300 4-chlorobiphenyl 198, 303, 304, 305,

438 metabolism 300, 301

3-chlorocatechol 192, 199, 200 4-chlorocatechol 199 Chloroform 3, 4, 102, 113, 118, 130,

256, 268, 343, 349, 364, 365, 366, 396, 421, 425, 444, 446, 447, 448, 488, 503

Chloroformic extraction 369, 370 Chlorograms 2

Chloroisopropylbiphenyls 28 4-chloro-4'-isopropylbiphenyl 29 Chloromethyl methyl ether 427 Chloronaphthalenes 211 metabolism 302

Chlorophenols 270, 430 Chlorophenoxyacetic acids 212-214 Chlorophyll 70, 71 p-chlorotoluene 193 Cholinesterase inhibitors 5, 270-272 Chromium 132, 136, 139, 140, 346,

421, 422, 423, 475-478 hexavalent 475-478

Chrysene 247, 427, 436 CI Acid Blue 1 156-157 Classification

functional group system 40 of carcinogens 419, 424, 431 of pollutants 7, 8, 39, 43, 96

Clay 467, 468 Closed-loop gaseous stripping 103,

104, 108, 109, 112, 115, 363 Cobalt 422, 423

Cocarcinogens 127, 423, 430, 450 Coho salmon 351, 352 Cometabolism 190 Commission of the European Communities

(CEC) 33, 40, 41, 43, 330, 331 list of contaminants 431

Common Market — see European Economic Community

Concentration factor 289 Constance, Lake 128 Control of Pollution Act (GB) 29 Co-oxidation 189-190 Copper 139, 140, 422, 423 COST Project 64b 38, 40, 41, 43, 87,

121, 149, 387-388 Council of Europe 324 Cramer method 23 Crawford, A. 299 Cresols 150 m-cresol 196 p-cresol 196 Crosby pesticide hazard rating 17, 22,

23 Cumene 223, 226, 228-235 Cunners 353 Cunninghamella elegans 193, 195 Cyanide 136, 139, 140, 343, 344, 503 4-cyanovaleric acid 228, 232 Cycloalkanes 72, 74, 82 Cyprinodon variegatus 314 Cyprinus carpis 481 Cytochrome P-448 484 P-450 299

Cytotoxicity 370, 372, 373, 378, 379, 380, 409

assays, in vitro 370

Daphniae 2, 5 Data banks 33, 34, 39, 40 DDE 209, 242, 246, 247, 254, 255,

256-257, 427, 444 DDT 2, 4, 23, 98, 149, 150, 187, 207,

208, 209, 239, 247, 253, 254, 420, 427, 444, 481, 494

de Geer, D.J. 323 de Greef, E. 495 Deaminases 299 Decanal 500 n-decanal 110 Degradable chemicals 275-280 Degradation

of aromatic pollutants 197 of chlorinated benzoic acids 199-201 of chlorinated biphenyls 197-199

DEHP 481 Demethylases, o- and n- 299 Demineralization 399 DENA 4

512 Index

den Boer, A. 499 Dermal contact 1 Derwent, River 276 Detergents, anionic 278-279 Detoxification 299, 302-303 Diagenesis 74, 80 Diazinon 247 Dibenz(a,h)anthracene 425 1,2,3,4-dibenzanthracene 484 Dibenzo-p-dioxins, chlorinated 211 Dibenzothiophene 242 Dibromochloromethane 446 Dichloroaniline 273 3,4-dichloroaniline 207, 272, 366 Dichlorobenzene 103, 113, 269, 366 m-dichlorobenzene 343 o-dichlorobenzene 343 1,2,4-dichlorobenzene 102 1,4-dichlorobenzene 101, 102, 103,

110, 112, 113, 114, 116, 118, 119, 121

3,3'-dichlorobenzidine (DCB) 473-474 photodegradation 473-474 microbial degradation 474 metabolism 474

3,4-dichlorobenzoic acid 193 3,5-dichlorobenzoic acid 193 Dichlorobiphenyl 287 4,4'-dichlorobiphenyl 198, 300 Dichlorobromomethane 446 1,2-dichloroethylene 101, 102, 110,

111, 112, 121 2,5-dichloro-4'-isopropylbiphenyl 26,

27 2,4-dichlorophenol 201, 247, 430 Dichloropyridines 25 Dieldrin 4, 205, 207, 251, 427, 444 Diethyl phthalates 150 Diffusivity 177, 179 Dihydrodiols 187, 192, 193 Dihydroxybenzenes 25 l,2-dihydroxy-l,2-dihydronaphthalene

193 Dihydroxynaphthalenes 25 Dirnethoate 271 1,1-dimethoxyisobutane 499 2,4-dimethylaniline 428 3,5-dimethylbenzoic acid 193 4,5-dimethylcatechol 189 2,5-dimethylfuran (DMF) 248, 250 1,4-dioxane 425 Dioxetane 225 Dioxins, chlorinated 214 Diphenylamine 209 Diphenyl ethers, chlorinated 218 Diphenylhydrazine 444 1,2-diphenylhydrazine 428 Diphenylmercury 242 Discriminant analysis 23

Disinfection 365 Dispersion tendency 20 Distribution of organic pollutants 39,

40, 45, 46-54, 56, 58-62 Distribution Register of Organic Pollu-

tants in Water — see WaterDROP Disulfotan 250 Diterpanes 71 DMAS 4 DMN 4 DNA 448, 482 DNBP 481 Dogfish shark 483, 484 Dove, River 276 Drinking water 1, 40, 55, 63, 101, 103,

104, 108, 114, 121, 130, 265, 266, 268, 269, 273, 330, 359-367, 370, 372, 373, 396, 401, 402, 419, 421, 495, 501, 503

colour 359 odour 359, 362, 366, 367 organic compounds in 387-388, 405,

407, 443 rate of consumption 360, 361 sensory assessment 361-362 standards 397-398 taste 359-367, 500 temperature 356 testing 405 turbidity 359

ECDIN 33, 34, 35, 37, 38, 43 Eco-Core microcosm 315-317 Ecokinetics 13, 15, 16 Ecological limit 6, 7 Ecosystems 309, 311, 318, 319

estuarine 310, 314, 318 laboratory models 493 salt-marsh 317

EDTA 154-155, 242 Effluents 40, 57, 63, 101, 114, 119,

120, 135-140, 144, 276, 385, 386, 503

industrial 135-140, 142, 144, 145, 146-148, 149, 152

recycling of 383-392, 395-402 registration of 171 sewage 143, 152, 155 treatment of 402

Eglinton, G. 69 El-Barbary, Ibrahim 467 Electron transfer processes 207, 209 Emission

registration 171 standards 331

Endosulfan 5 Energy transfer 245-246 Environment Contaminants Act (Canada)

29

Index 513

Environmental Chemicals Data and Infor-mation Network — see ECDIN

Environmental chemodynamics — see Eco-kinetics

Environmental Mutagen Information Cen-ter (EMIC) 432

Environmental policy European 323, 330, 332 international 332

Environmental processes 74, 80, 224, 227

Environmental Protection Agency (EPA) 18, 39, 40, 63, 64, 309, 396, 419, 420, 432, 447

Environmental Research Programmes 33 Enzymatic oxygen fixation 189 Enzyme induction 353 Enzyme systems 197 Enzymes 299 Epichlorohydrin 428, 436 Epidemiological evaluation 399-400 Epidemiological studies 350, 390-391,

446-447, 450 Epizootiological studies 351 Esters

phosphate 97 phosphoric 270, 272 phthalate 96 phthalic acid 481

Estuarine ecosystems 310, 314, 318 Ethers 98, 499

chlorinated 96-97 Ethylbenzene 110, 178, 192, 193, 365 Ethyl carbamate 425 Ethylene dibromide 425, 434 Ethylenethiourea 244, 426, 434, 444 4-e thy1-su1phonylnaph tha1ene-1-su1fon-

ami de 3 Ethynyl oestradiol 156-157 EURONET 38 European Economic Community (EEC) 29,

174, 327, 329, 330 Evaluation, epidemiological 399-400 Exhaust, vehicle 487, 488

Fatty acids 96 Ferrand, R. 87 Fertilizers 137 Fick's law 179, 181 Field studies 238 Filtration 495

activated carbon 268, 271, 363 bank 265-273, 362, 499

Fish 315, 331 breathing patterns 343 gill lesions 346 human consumption of 1 respiration monitoring 501-503 studies for tumours 350-351, 449

Fish (cont'd) toxicity monitoring 342, 343, 471

Flow-through tanks 341 Fluctuation assay procedure 389 Fluoridation 402 Fluoride 422 3-fluorobenzoate 193 4-fluorobenzoate 193 p-fluorotoluene 193 Food chains 2, 311 Foureman, Gary L. 483 Fox River watershed 449 Frameshift mutagens 416 Free-energy constants 21 Free radicals 224, 226, 228, 235, 237,

247, 248, 251 Functional group classification system

40 Fundulus grandis 314-315

Gab, Siegmar 465 Garland, J.H.N. 275 Garrison, Arthur W. 39 Gas chromatography 40, 64, 65, 70, 87,

88-89, 94, 95, 103, 104-105, 109, 110, 113, 288, 461, 487, 488, 494

Gaseous stripping 103, 104, 108, 109, 112, 115, 363

GC-MS analysis 55, 64, 65, 66, 67, 70, 75, 87, 89, 92, 94, 105, 110, 131, 156, 157, 363, 364, 390, 405, 487, 500

Gebefugi, Istvan 465 Geiss, F. 33 Gentisic acid 196 Geosmin 500 Gibson, David T. 187 Giger, Walter 101 Glatt River 103, 112, 113, 120, 472 Goiterogens 351 Grass shrimp 314 Great Lakes 350-351 Grey list of pollutants 7, 8, 17, 327,

328, 331, 367, 507 Groundwater 101, 103, 105, 106, 107,

108, 113, 118, 121, 128, 364, 499 as source of drinking water 265 replenishment 266

Gujer, W. 471 Gulf killifish 314

H-atom transfer 231, 247, 248 Haloforms 132, 150, 364, 367, 496 Halogen 215 Halogenated compounds 4 Halogenated hydrocarbons 207

volatile 102 Halonaphthalenes 211 Hansch model 17, 21, 23

514 Index

Hazard rating systems 17, 22, 23 Health & Safety at Work Act (GB) 29 Health hazards 369, 383

of wastewater recycling 395 Heavy metals 467-468 Hela cells 370, 378, 379 Hela test 370, 372, 373, 379 Helmes, C. Tucker 419 Hemon, Denis 369 Hendry, Dale G. 223 Henry's Law constant 175, 176, 181-182

183 Hepatocarcinogens 353 Hepatotoxic compounds 346 Heptachlor 428, 444 Heptachlor epoxide 428, 444 Herbicides 201, 207, 212, 215, 218 Hexabromobenzene 25, 288 Hexabromobiphenyls 25 Hexachlorobenzene 25, 247, 285, 494 Hexachlorobiphenyls 25, 28, 494 Hogendoorn-Roozemond, A.S. 475, 477 Hollod, Gregory J. 467 Hopanes 75, 78, 79 Humic material 251 Hutzinger, 0. 13 Hydrases 299 Hydrocarbons 69, 71, 73, 74, 87, 96,

97, 189, 287, 264, 365 aliphatic 103 alkylated aromatic 110, 120 aromatic 103, 187, 188, 192, 193 chlorinated 103, 104, 126, 461, 465,

481 chlorinated aromatic 300-302 halogenated 207 petroleum-derived 110 polyaromatic 343 polycyclic 188, 192 polycyclic aromatic 126, 128-129,

132 volatile chlorinated 101-212, 130,

132 volatile halogenated 102

Hydrolases 299 Hydroperoxide 225, 231, 235, 248 Hydrophobie fragmental constants 21 Hydrophobie pollutants 177 m-hydroxybenzoic acid 196 p-hydroxybenzoic acid 196 Hydroxylases 299, 483, 484 Hydroxylation 189, 191, 300, 303 p-hydroxymandelic acid 196 4-hydroxy-l-tetralone 193

Immunosuppression 3 Indane 110 Indeno(l,2,3-c,d)pyrene 426 Industrial effluents 135-140, 142, 144

Industrial effluents (cont'd) 145, 146-148, 149, 152

Injection wells 266 Insecticides 217, 218 Interfacial processes 176, 177 International Reference Centre for

Community Water Supply 43, 384 Iron 422, 423 Isoprenoids 75, 78, 79, 81 Isopropyl phenols 228, 231-232, 235

9

James, Margaret 0. 483

Keith, Lawrence H. 39 Kepone 428, 444 Ketones 98, 103, 499 Killifish 314 Kinae, Naohide 481 Kinetic analysis 228 Knoeppel, H. 487 Koeman, J.H. 341, 475, 477 Kohli, J. 299 Kolar, Z. 475, 477 Kopfler, F.C. 405 Kraybill, H.F. 419 Kussmaul, Horst 265

Laboratory model ecosystems 493 Lake surface water 487 Lang, D.R. 405 Lazar, Philippe 369 Leaching 20 Lead 139, 140, 421, 422, 423 Leaf wax alkanes 71 Lee, River 153, 154-155, 156, 157, 385,

395 Lek, River 479 Light attenuation 252 Limmat River 106, 107, 112 Lindane 23, 131, 428, 444, 503 Lipids 69, 493 Lipophilicity 26, 28 Liquid chromatography 66-67, 88, 95 Little skates 483, 484 Locomotor response of fish 342 Loper, J.C. 405 Lugworm 312-313 Lunde, Gulbrand 461

Maas, River 139 Mackay, Donald 175 Maier, Walter J. 463 Main River 269 Malathion 251, 271 Mandelic acid 196 Manganese 422, 423 MARC reports 17 Mass spectrometry 40, 65, 66, 88, 92

, Mass transfer 176, 180-181, 183

Index 515

Mathematical models 17, 22, 237, 317, 464

Maxwell, J.R. 69 Mazza, M. 87 McDuffie, Bruce 467 Mercuric oxide 285 Mercury 131, 136, 137, 139, 140, 179,

422, 481, 507 Metabolism microbial 196 of aromatic compounds 187, 188-190 of chlorinated aromatic hydrocarbons

300-302 of chlorinated benzene 302 of PCBs 300-302 of PCNs 302 of xenobiotics 299 studies 317

Metabolites, toxic 341 Metalloids, toxic 132, 133 Metals 421, 422, 423, 461

heavy 467-468 toxic 131, 133 trace 467

Methoxychlor 209, 210, 243, 247 3-methylbenzoic acid 193 4-methylbenzoic acid 193 3-methylcholanthrene (MCA) 411-412,

484 Methyl iodide 428, 436 2-methylisoborneol 500 Methylisobutanoate 499 Methylmercuric chloride 242 Methyl methanesulfonate 411 Methyl parathion 242, 311, 313, 314,

315, 316, 317 a-methylstyrene 193 Meuse, River 507 Microbial metabolism 196 oxidation 193 transformations 187-201

Microcosms 309, 312, 317-318 Micropollutants 449 Microsomal hydroxylase 483, 484, 485 Microtracers 101 Mill, Theodore 223 Mineral oil 136, 139, 140 Mirex 428 Mississippi River water 349, 395, 396 Modelling mathematical 17, 22, 237, 317, 464 of aquatic processes 183 of ecosystems 493

Molnar-Kubica, Eva 101 Molybdenum 422 Morra, C.F. 359, 499 Muconic acid 189 Mudminnow 345

Multidetection analysis 87 Mutagenesis 405, 412, 413

testing 408, 409 Mutagenicity 2, 8

tests 345, 389-390, 433, 481-482 Mutagens 432-440, 441, 450

frameshift 416 identification of 431-432

NADPH cytochrome-C reductase 299 Naphthacene 247, 250 ß-naphthalamine 4 Naphthalene 110, 178, 181, 192, 193,

195, 196, 210, 247, 365 chlorinated, metabolism 302 derivatives 24-25

α-naphthoflavone (a-NF) 483, 484, 485 a-naphthol 193 0-naphthol 193 1-naphthol 247 1,4-naphthoquinone 193 2-naphthylamine 426, 434 Neoplasms in aquatic organisms 350-351,

481 Nephrotoxic compounds 346 New Orleans study 396 Nibea mitsukurii 481-482 Nickel 139, 140, 421, 422, 423 NIH shift 301, 302 Nitrate 471 Nitrification 279, 280, 471-472 Nitrilotriacetic acid (NTA) 242, 428 Nitrite 471, 472 Nitrobacter 471 Nitrobenzene 247 Nitrobenzoic acid, m- and p- 196 4-nitrobiphenyl 426, 434 Nitrodiphenyl ether 218 Nitrofen 218 p-nitrophenol 247 Nitropropane 428 4-nitroquinoline-n-oxide 411 Nitrosamines 3, 126, 131, 132, 507 n-nitrosoatrazine 242 Nitrosomonas 471, 472 n-nonanal 110 Non-biodegradable matter 135-140 Non-chloroform trihalomethanes (NCTHM)

446, 447 North Sea 73, 325 Nucleophilic displacement 243

Octanol/water partition coefficient 24, 25, 28

Odour threshold concentration 500 Ofstad, Elizabeth Baumann 461 Ohio River 395, 446 Oil pollution 69, 70, 73, 75, 78, 79,

136, 139, 140, 169, 420

516 Index

Okefenokee Swamp 244, 246, 249 Oncorhynchus kisutch 351 Organic pollutants 39-40, 144-146,

160-168, 487 analysis 39, 40, 55, 63-67 carcinogenic 421, 423 chlorinated 461 classification 39, 43 distribution 39, 40, 45, 46-54, 56,

58-62 in drinking water 387-388, 405, 407,

443 occurrence 39, 40 persistence 150, 152 US national survey 447

Organochlorine compounds 268-269, 359 Organofluorine compounds 269-270 Organohalides 396, 496 Organophosphate pesticides 314 Organophosphorus compounds 345, 346 Osmosis, reverse 405, 406, 413, 416 Ott, H. 33 Oxidases 299 Oxidation 223-235, 278, 279

biochemical 463-464 biological 471 environmental 224, 227 free radical 251 microbial 193 of sulphides 250 photosensitized 248-251

Oxygen 208, 247 biological demand 276-280, 285 singlet 224, 237, 246, 248, 249,

250, 251 Oxygen fixation, enzymatic 189 Ozonization 130, 133, 265, 268, 271,

361, 495-496 Ozonolysis 150

Pack, E. 473 Packham, R.F. 383 Palaemonetes pugio 314-315 Papillomas 351 Paraffin hydrocarbons 96 Parathion 23, 218, 242, 244, 246, 247,

271, 343 Parlar, Harun 465 Particle size 290 Partition coefficients 20, 21, 24 Partitioning 253-255 Payen, P. 87 PCNB 445 Pentacene 250 Pentachloronitrobenzene 428 Pentachlorophenol 214, 242, 343 Pentachloropyridine 25 Peroxides 226, 251 Peroxy radicals 224, 230, 235

Persistence 5, 6, 7, 8, 14, 16, 19, 493

of organic pollutants 150, 152 testing 18

Pesticides 23, 55, 137, 205, 206, 242, 243, 244, 249, 270, 314

chlorinated 2, 269 hazard ratings 17, 22, 23 photochemistry 242

Petroleum-derived hydrocarbons 110 Phenanthrene 178, 192, 194, 196 Phenols 24, 98, 149, 193, 196, 218,

247, 278-279, 343, 344, 346, 430 chlorinated 98

Phenoxyacetic acid 201 Phenylcarbamate 217 Phenyl ketones 247 Phenylmercurie acetate 242 Phenylurea 215 Philp, R.P. 69 Phosphate esters 97 Phosphatidyl choline 299 Phosphoric esters 270, 272 Phosphorous pesticides 270 Photochemistry 205-219 Photodegradation 20, 205, 209-212, 215 Photolysis 207-218, 237-257

direct 237, 239-244 prediction of 237-239 rates 239, 240, 257 sensitized 237, 244-251

Photosensitization 224 Photosensitizers 237, 244 Phototransformation 205-206 Phthalate esters 96 Phthalates 103, 461 Phthalic acid 196

esters 481 Phytoplankton 71, 75, 128, 238 Picloram 215 Piet, G.J. 359, 499 Poels, C.L.M. 341, 479 Pollutant stress 309 Pollutant wash out 488 Pollution abatement 169-174 Pollution of Surface Waters Act 135 Pollution potential 13, 16 Pollution profiles 461 Polychlorinated biphenyls (PCBs) 97,

126, 130-131, 132, 179, 181, 187, 243, 283, 285, 287-288, 352, 428, 445, 461

metabolism 300-301 photodegradation 210-211

Polychlorinated naphthalenes (PCNs) 285 metabolism 302

Polychlorinated terphenyls (PCTs) 212 Polycyclic aromatic hydrocarbons (PAHs)

98, 126, 128, 129, 132, 343, 363,

Index 517

Polycyclic aromatic hydrocarbons (PAHs) (cont'd) 420, 484

photochemistry 242, 250 Polycyclic hydrocarbons 188, 192 Polymers 288-289 Polystyrene 289 Polyvinylchloride 289 Porphyra tenera 352 Potassium permanganate 267 Precipitations 128 Predator-prey relationships 314 Pristane 70, 71, 75, 76 Propylene oxide 428 Protocatechuic acid 196 Pseudomonas 190, 191, 193

P. fluorescens 199 P. putida 192, 193

Pseudopleuronectes americanus 483, 485 Pyrene 247, 440 Pyrolysis 268, 269

QSAR 17, 21, 23, 25 Quantum yields 211, 213, 237, 242, 243,

244 Quenchers, triplet 208 Quenching 243 Quinic acid 196 Quinoline 242, 247, 428 Quinones 97

Radical intermediates 207 Radicals

alkoxy 224, 230 free 224, 226, 228, 235, 237 peroxy 224, 230, 235

Rainbow trout 353, 475-480, 501, 503 Rainfall 120, 463 Raja erinacea 483 Reaeration 175, 179 Recycling of effluent 383-392, 395-

402 Registration

of effluents 171 of pollutants 39

Reservoirs, open 499, 500 Residual organics 405, 407 Resorcinol 25 Reverse osmosis 405, 406, 413, 416 Rheotaxis 342, 343 Rhine, River 5, 7, 73, 112, 113, 128,

129, 131, 139, 169-174, 265, 266, 267, 269, 270, 272, 323, 325-329, 345-346, 361, 395, 479-480, 495, 499, 507

Rhine Chemical Convention 327-328, 329, 367, 507

Rhodamine B 428 Riboflavin 246, 247 Richardson, Harold 223

Ring fission 189, 190-196 Risk assessments, human 450 River water 96, 142, 373, 383, 385,

467-468, 499 Rotary-flow technique 343 Rotenone 247 Ruzo, L. Octavio 205

Saccharin 4, 428 Safe, S. 299 Safe Drinking Water Act (1974) 419 Safe substances 295 Safrole 426, 435 Saitou, Shinichi 481 Salamanders, neoplasia in 351-352 Salmo gairdnerii 345, 475-480, 501 Salmon 24, 351, 352 Salmonella/microsome system assay 405,

408, 409 Salmonella typhimurium TA 1538 481 Salt-marsh ecosystems 317 Sand 467, 468 Santema, P. 507 Sasaki, S. 283 Schoeny, R.S. 405 Sediments 20, 72, 75, 76, 77, 179, 253,

254, 316, 467, 468 Selenium 132, 421 Sewage 142, 143, 152, 155, 383

treatment 150 Shackleford, Walter M. 39 Sheepshead 483, 484 Sheepshead minnow 314 Shellfish, disease in 351 Shikimic acid 196 Shuval, Hillel I. 395 Sigman, Caroline C. 419 Sihl River 106, 107 Sikka, H. 473 Silicones 96 Silt 467, 468 Simazine 428, 445 Singlet oxygen 224, 237, 246, 248, 249,

250, 251 Skin tumours 351 Slooff, W. 341, 501 Sludge, activated 29, 150, 285 Smith, C.C. 405 Snow 487 Sodergren, Anders 493 Soils 72 Solubility 286 Sonstegard, R.A. 349 Squalus acanthias 483 Stabilities, chemical 20 Standards drinking water 397-398 effluent 331 surface water 5

518 Index

Standards (cont'd) wastewater reuse 397 water quality 172

Statistical studies 446 Steranes 71, 72, 74, 75, 76, 78, 79,

82 Stereochemistry 74, 75, 76 Steroids 98 Stichting Europoort Botlek Belangen

170, 174 Storms, effect of 468 Stress, pollutant 309 Strik, J.J.T.W.A. 475, 477 Stripping techniques 390 Structure-activity relations 17, 21 Styrene 149, 288-289 Sulphides, oxidation of 250 Sulphur heterocyclic compounds 98 Sundstrom, Goran 205 Superior, Lake 463-464 Surface films 182, 255, 461, 493, 494 Surface runoff 128 Surface water 40, 57, 141, 152, 171,

364, 487, 503 pollution 135 standards 5

Susquehanna river basin 467-468 Swain, Wayland R. 463

Tame, River 276 Tannic acid 429 Tap water 128, 360, 363, 499

analysis 103 Tardiff, R.G. 405 Taste of drinking water 359-367, 500 ten Holder, V.J.H.M. 475, 477 Terpenes 157 terphenyls 210 polychlorinated 212

3,3' ,4,4'-tetrachloroazobenzene 207 1,2,3,4-tetrachlorobenzene, metabolism

302 2,2f,5,5'-tetrachlorobiphenyl 300 2,3,7,8-tetrachlorodibenzo-p-dioxin

(TCDD) 212, 214 photodegradation 465-466

1,1,2,2-tetrachloroethane 437 Tetrachloroethylene 101-121, 268, 488 1,1,2,2-tetrachloroethylene 429 Tetrachloromethane 365 Tetrachloropyridines 25 Thames, River 385, 395 Thermoclines 177, 179 Thyroid hyperplasia 351 Tiberio, Robert D. 467 Tin 422 Toluene 110, 119, 120, 192, 193, 196,

343, 365, 487 p-toluic acid 196

Toluidines 367 o-toluidine 426 p-toluidine 429 Tomita, Isao 481 Total organic carbon (TOC) measurements

66, 463, 464 Town, W.G. 33 Toxaphene 346, 429 Toxic Substances Control Act (TOSCA) 29 Toxic substances laws 14, 17, 29 Toxicity 2, 6, 7, 8, 13, 14, 16, 35,

342, 409 chronic 287 monitoring 342 testing 17-19

Toxicology 389, 398-399 data 388

Toxification 299 Trace metal concentrations 467 Transferases 299 Transformation assay, mammalian cell

405, 408, 414, 415 Transformations, microbial 187-201 Transport 20, 468 Trent, River 275-280, 386 Triazine 215, 216, 248 Trichlorobenzene 103, 269, 288, 366 1,2,4-trichlorobenzene 110, 112, 114,

116, 118, 119, 121 Trichloroethane 113 1,1,1-trichloroethane 101, 102, 108,

110, 111, 112, 114, 116, 118, 120, 121, 439, 488

1,1,2-trichloroethane 429 1,1,2,2-trichloroethane 429 Trichloroethylene 4, 101, 102, 110, 111,

112, 113, 114, 116, 118, 120, 121, 149, 268, 343, 437, 445, 488

1,1,2-trichloroethylene 429 2,4,5-trichlorophenoxyacetic acid 429 Trichloropyridines 25 Trifluoromethylaniline 273 3-trifluoromethylaniline 272 4-trifluoromethylbenzoate 193 Trifluralin 242 Triglycerides 98 Triplet diradicals 224 Triplet quenchers 208 Triplet state energy 245, 246, 247 Triterpanes 71, 72, 74, 76, 78, 79, 82 Trout 342, 345, 346

rainbow 353, 475-480, 501, 503 Tryptophan 196, 250 Tulp, M.T.M. 13 Tumorigenic compounds 346 Turbulence 179

Ultraviolet light 252 Umbra limi 345

Index 519

Umbra pymaea 345 UN Conference on Water Resources (1976)

1 United Nations Environment Programme

(UNEP) 8 US Tariff Commission reports 17

van de Velde, 0. 135 van Esch, G.J. 1 van Lookeren Campagne, N. 169 Vanadium 422 Vanillic acid 196 Vapour pressure 175 Versino, B. 487 Vinylchloride 3, 4, 126, 130, 256,

289, 290, 350, 426, 434, 444 Vinylidene chloride 429, 436 Vissers, H. 487 Volatile chlorinated hydrocarbons 101-

121, 130, 132 Volatile halogenated hydrocarbons 102 Volatile substances 388 Volatilization 175-184, 253-254, 256

Waggott, A. 141 Wakeham, Stuart 101 Wastewater, reuse of 383-392, 395-402 monitoring 400-401 standards 397

WaterDROP 39, 41, 42, 43 Water quality 323-324, 360, 463

control 400-401, 501, 503 criteria 309 European regulations 323 standards 172

Water Research Centre (UK) 40, 149, 151, 153, 157, 387, 390, 391, 392

Water resources, development 392 Water supplies 105

sources 385 Water treatment 266-267, 268, 271, 373,

380, 384, 402 Weathering 72, 75 Weber pesticide hazard rating 17, 22,

23 Wheatland, A.B. 141 White list of pollutants 7, 8 White suckers 351 Whitman two-film model 175, 176 Wilkes, Frank G. 309 Windspeed 180 Winter flounder 483, 484, 485 World Health Organization (WHO) 43,

325, 399 Wyndham, C. 299

XAD resins 363, 390, 416 XAD-2 156, 405

Xenobiotics 205, 299, 319, 352 metabolism 299

Xylene 149, 343, 365 m-xylene 110 o-xylene 110 p-xylene 110, 193

Zebra fish 346, 502, 503 Zepp, Richard G. 237 Zinc 132, 136, 139, 140, 343, 422, 423 Zitko, V. 13 Zoeteman, B.C.J. 359, 499 Zooplankton 70, 71, 77 Zurich, Lake 101, 103, 104, 106, 114,

115, 116, 117, 118, 119, 120, 121, 131

aquatic pollutants. transformation and biological effects

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