Preface

The present thesis has been prepared at the Department of Environmental Science and

Engineering, Technical University of Denmark, as part of the fulfilment of the Ph.D.

degree requirements. The work was carried out in the period from March 1997 to Au-

gust 2000, with Associate Professor Hans-Jørgen Albrechtsen as supervisor and As-

sociate Professor Hans Mosbæk as co-supervisor. The thesis consists of a summary

focused on pesticide degradability in groundwater, with specific emphasis on the

possible role of the natural redox environment of aquifers, and the following five

papers of which the first four are submitted to international peer-reviewed journals

(not included in www version).

I: Pedersen PG, Albrechtsen H-J (2000a) Spatial variability of redox processes within

and between seven shallow aquifers in Denmark.

II: Pedersen PG, Albrechtsen H-J (2000b) Microbial activity in aerobic and anaerobic

aquifers assessed by the turn-over potential of benzoic acid.

III: Pedersen PG, Mosbæk H, Albrechtsen H-J (2000) Fate of eleven pesticides:

Degradability and sorption in eight Danish aerobic and anaerobic aquifers.

IV: Arildskov NP, Pedersen PG, Albrechtsen H-J (2000) Fate of the herbicides 2,4,5-

T, atrazine, and DNOC in a shallow, anaerobic aquifer investigated by in situ passive

diffusive emitters and laboratory batch experiments. Accepted.

V: Pedersen PG, Albrechtsen H-J (2000c) Slow sorption of atrazine in an anaerobic

low organic carbon aquifer. In preparation.

Since one of the purposes of the summary is to compare the above mentioned papers

with the litterature, all references to the papers are accentuated with bold roman

numerals (I-V).

Many people were involved in this project, and it could not have been fulfilled

without them. Employees at the Department of Chemistry, Royal Veterinary and

Agricultural University, Copenhagen, treated me nice during my visit there. Associate

Professor Christian Bender Koch should be especially mentioned. The people wor-

king at my ”own” department were always professional and ambitious in a pleasant

way, providing a strong daily day basis for the work. A large number of people, both

at the department and other institutions, performed different analysis for me once in a

while. On a more regular basis Carina Aistrup and Mona Refstrup took samples and

generated high quality data for me to work on, and they should be especially thanked.

Torben Dolin and Birte Brejl made beautiful drawings, and Grete Hansen and Helle

Offenberg provided litterature. During field trips, Mads Georg Møller, Haraldur

Hannesson, Bent Skov and Jens Schaarup Sørensen, did the hard work. My co-

supervisor, Hans Mosbæk, often did miracles to a sick, non-functioning HPLC. My

supervisor, Hans-Jørgen Albrechtsen, was always prepared to read and thoroughly

comment manuscripts, and to produce new ideas for me to work on.

The last thanks to my family, for being there.

Philip Grinder Pedersen

ii

Abstract

The presence in groundwater of xenobiotic compounds like pesticides causes concern

since groundwater in many countries is used for drinking water purposes and indu-

strial and agricultural use, and since natural ecosystems may be fed by groundwater

containing pesticides. However, investigations on pesticide fate in groundwater are

few in number, especially considering the large number of pesticides used in society

and the large number of pesticides detected in groundwater.

This Ph.D. thesis presents and discusses current knowledge on pesticide degradability

in groundwater. Since many pesticide transformation processes may only be possible

under certain redox conditions, special emphasis is given to the possible influencing

role of the natural aquifer redox conditions. Sources of pesticide pollution and

ecotoxicological and health aspects of groundwater contamination with pesticides are

not dealt with.

The experimental part of the Ph.D. project focused on two interlinked topics. One was

to characterize the natural redox conditions and other conditions of a number of

different pristine aquifers, mainly anaerobic. The other was to investigate the degrada-

bility behaviour of a number of pesticides in samples from the different aquifers, and

eventually to couple the behaviour to the aquifer redox conditions or other charac-

teristics.

Point source contaminations – e.g. with chlorinated solvents, oil or landfill leachate –

may alter the natural redox conditions of an aquifer towards a more reduced environ-

ment, whereas pesticide pollutions typically are low load pulse pollutions, which

probably does not alter the natural redox conditions, except in extreme cases.

Therefore the ”natural” redox environment of aquifers is relevant to consider when

assessing the fate of pesticides in groundwater.

iii

Natural aquifer redox conditions were shown to vary substantially between and within

the investigated aquifers. Aerobic conditions were not always present and redox

conditions varied within a few meters of vertical distance in some aquifers. The redox

environment of aquifers is a function of several factors, including geological origin of

the aquifer settings, flow patterns, and organic load. Without detailed investigations it

may be difficult to assess the natural redox environment of a pesticide polluted

aquifer.

The coupling between pesticide degradability and anaerobic conditions has been

studied in a few investigations only, typically with a single pesticide and a specific

aquifer redox environment. In the experimental part of this project, using a range of

pesticides and a range of aquifers and aquifer redox environments made it possible to

both expand the number of pesticide investigations in anaerobic aquifers and to

compare pesticide behaviour between aquifers. Since the experimental setup was

similar in all the performed experiments, the found differences in pesticide behaviour

were not due to setup differences but could be ascribed differences in aquifer charac-

teristics.

Pesticides are a chemically very diverse group of compounds, and the degradability is

equally diverse. Some pesticides – like the phenoxy alcanoic acid (e.g. MCPP) or s-

triazine (atrazine) pesticides – have been subject to a relatively large number of

groundwater investigations, including the ones performed in this project. General

trends can be extracted from the investigations but ”conflicting” behaviour is often

observed. Several investigations documented phenoxy alcanoic acid degradation in

groundwater under mainly aerobic conditions, but a few showed persistence under

both aerobic and anaerobic conditions. Atrazine was persistent under both aerobic and

anaerobic conditions in some investigations, whereas others showed that atrazine was

degradable or at least removed from the groundwater phase under anaerobic condi-

tions. Clearly, the mechanisms controlling both atrazine and phenoxy alcanoic acid

degradability in groundwater are not completely understood, and may be a function of

aquifer specific factors as well as type of pesticide.

Other pesticides – e.g. the acetanilide (e.g. alachlor) or phenylurea (e.g. isoproturon)

pesticides – probably will show the same complex behaviour if more investigations

iv

are made. None the less, additional investigations of the behaviour of these pesticides

are needed to better understand the fate of these pesticides in groundwater.

It may be operationally and regulatory sound to neglect the importance of ground-

water degradability when evaluating the fate of pesticides in the environment. From a

precautionary principle viewpoint this is a responsible approach, especially since the

experiments performed in this project and other investigations showed that some

pesticides, e.g. bentazone and the dichlobenil metabolite BAM (2,6-dichloroben-

zamide) seemed to be persistent or very slowly degradable in groundwater, regardless

redox or other conditions. The variable degradation behaviour between aquifers of the

degradable pesticides also shows that general assessments of pesticide degradability

and degradation rates should be avoided. Use of pesticides should not be allowed

based on assumed groundwater degradability only.

On the other hand, degradation of pesticides in groundwater might in certain cases be

the last possible way of minimizing a pollution problem, making a continued effort to

understand the degradation processes in aquifers relevant. It is especially important to

document whether the degradability observed in controlled and to a certain extent

“artificial” degradation experiments with high and constant pesticide concentrations

can be transferred to “real world” behaviour. However, since many factors influence

the presence of pesticides in groundwater, it may be difficult in “real world” systems

to determine the importance of aquifer degradability in comparison to e.g. top soil

attenuation processes.

In order to better understand the role of the aquifer two approaches can be proposed.

One approach is to exploit the presence of pesticide metabolites in groundwater in

comparison to parent compound (e.g. using metabolite/parent compound ratios). Low

ratios would show that only minimal degradation had occurred, both in top soil and

groundwater. High ratios would in itself only show that the parent compound had

been subject to degradation, either in the aquifer or prior to being transported to the

aquifer. However, combined with e.g. groundwater age determinations it might be

possible to indicate degradation in groundwater.

v

Another approach is to integrate groundwater characteristics in “real world”

investigations on pesticide detections. Groundwater characteristics would include

redox conditions (e.g. presence and concentration of redox relevant compounds, like

oxygen or iron) or information on previous exposure to pesticides. Exploiting such

knowledge in e.g. statistical analysis on groundwater detections might in some cases

lead to detection of degradability “trends”, e.g. lack of detecting certain pesticides

under specific redox conditions. Such trends, indicative of groundwater degradation

or non-degradation under specific groundwater conditions, could subsequently be

investigated and eventually verified in more controlled investigations.

vi

Resumé

Fund af pesticider og andre miljøfremmede stoffer i grundvand giver anledning til

bekymring, både fordi grundvand i mange lande benyttes som drikkevand, industrielt

og i landbrug, og fordi naturlige økosystemer i hydrologisk kontakt med grundvand

(fx vandløb og søer) kan blive påvirket af pesticidforurenet grundvand. Der er dog

kun lavet relativt få undersøgelser af pesticiders skæbne i grundvand, især taget i

betragtning det store antal pesticider, der bruges i samfundet og findes i grundvandet.

Denne ph.d.-afhandling præsenterer og diskuterer den nuværende viden om pesticid-

nedbrydelighed i grundvand. Mange nedbrydningsprocesser er kun mulige under

bestemte redoxforhold, og derfor er der lagt særlig vægt på det naturlige redoxmiljø i

grundvand. Kilder til forurening med pesticider og helbreds- og økotoxikologiske

aspekter er ikke behandlet.

Den eksperimentelle del af ph.d.-projektet fokuserede på to koblede emner. Det ene

var karakterisering af de naturlige redoxforhold og andre forhold i et antal forskellige

uforurenede og primært anaerobe grundvandsmagasiner. Det andet var at undersøge

nedbrydeligheden af et antal pesticider i prøver fra de forskellige magasiner, og om

muligt at koble nedbrydeligheden til magasinets redoxmiljø eller andre forhold.

Punktkildeforureninger – fx med chlorerede opløsningsmidler, olie eller losseplads-

perkolat – kan ændre det naturlige redoxmiljø mod mere reducerede forhold, mens

forureninger med pesticider typisk består af små mængder pesticid i pulse, som –

undtagen i særlige tilfælde – sandsynligvis ikke vil ændre det naturlige redoxmiljø.

Derfor er det naturlige redoxmiljø i grundvand vigtigt at tage med i betragtning, når

pesticiders skæbne i grundvand skal vurderes.

Det blev påvist at redoxforholdene i de undersøgte grundvandsmagasiner varierede,

både indenfor det enkelte grundvandsmagasin og mellem magasinerne. Således var

der ikke aerobe forhold i alle magasiner, og i nogle magasiner skiftede redoxmiljøet

vii

vertikalt indenfor få meter. Redoxforholdene styres af mange faktorer, bla grund-

vandsmagasinets geologiske dannelsesmiljø, strømningsforholdene og mængden af

organisk stof. Uden detaljerede undersøgelser kan det være vanskeligt at vurdere

redoxforholdene i et pesticidforurenet grundvandsmagasin.

Der er kun lavet få undersøgelser af sammenhængen mellem nedbrydelighed af

pesticider og anaerobe grundvandsforhold, og typisk kun med et fåtal af pesticider i et

specifikt redoxmiljø. I den eksperimentelle del af dette projekt blev mange forskellige

pesticider og mange forskellige grundvandsmagasiner med forskellige redoxmiljøer

undersøgt, hvilket både forøgede det totale antal grundvandsundersøgelser af

pesticider i forskellige anaerobe miljøer, og muliggjorde en sammenligning mellem

forskellige grundvandsforhold. De fundne forskelle i nedbrydelighed mellem prøver

fra forskellige grundvandsmagasiner skyldtes ikke forskelle i eksperimentelt setup,

idet disse var ens i alle eksperimenter, men kunne tillægges forhold relateret til grund-

vandsmagasinerne.

Pesticider er kemisk set meget forskellige, og der er ligeledes forskellig nedbrydelig-

hed af de forskellige pesticider. Visse pesticider – såsom phenoxysyrerne (fx MCPP)

eller s-triazinerne (fx atrazin) – har været genstand for et forholdsvist stort antal

undersøgelser, inklusive de eksperimentelle forsøg i dette projekt. Der kan uddrages

generelle tendenser af disse undersøgelser, men også “modstridende” opførsel. Aerob

nedbrydning af phenoxysyrer er dokumenteret i mange undersøgelser, mens nogle få

har vist manglende nedbrydning under både aerobe og anaerobe forhold. På samme

måde har et antal undersøgelser vist manglende nedbrydning af atrazin i grundvand,

både aerobt og anaerobt, mens andre har vist nedbrydning eller fjernelse fra

grundvandet. Der er således ikke en samlet forståelse af mekanismerne bag

nedbrydeligheden af atrazin eller phenoxysyrer i grundvand, men faktorer relateret til

det specifikke grundvandsmagasin har efter alt at dømme betydning, parallelt med de

stofspecifikke faktorer.

Andre typer pesticider – fx acetanilider (fx alachlor) eller phenylurea-pesticider (fx

isoproturon) – vil sandsynligvis opføre sig tilsvarende komplekst i nye undersøgelser

af grundvandsnedbrydeligheden. Dette viser nødvendigheden af sådanne under-

søgelser for at forbedre forståelsen af også disse pesticiders skæbne i grundvand.

viii

Det er operationelt og regulatorisk fornuftigt at se bort fra betydningen af grundvands-

nedbrydeligheden, når pesticiders skæbne i miljøet skal vurderes. Dette vil være

ansvarligt ud fra en forsigtighedsprincip-synsvinkel, idet resultaterne fra dette projekts

eksperimentelle del og andre undersøgelser har vist at visse pesticider – fx bentazon

og dichlobenilnedbrydningsproduktet BAM (2,6-dichlorbenzamid) – ser ud til at være

ikke-nedbrydelige eller kun meget langsomt nedbrydelige i grundvand, uanset redox-

forhold eller andre forhold. For de pesticider, som kan nedbrydes i grundvand, gør

forskelle mellem grundvandsmagasiner det også vanskeligt at give generelle

vurderinger af nedbrydeligheden. Brug af pesticider bør derfor ikke tillades på

baggrund af nedbrydelighed i grundvand alene.

Omvendt kan nedbrydning af pesticider i grundvandet være den sidste mulighed for at

minimere et forureningsproblem. I sådanne tilfælde er øget viden om nedbrydelighed i

grundvandsmagasiner derfor relevant. Det er særlig vigtigt at få dokumenteret,

hvorvidt den nedbrydelighed, der kan observeres i kontrollerede og til en vis grad

“kunstige” nedbrydningseksperimenter med høje og konstante pesticidkoncentra-

tioner, kan overføres til nedbrydelighed i virkelige grundvandsmagasiner. Denne do-

kumentation kan være vanskelig at få, fordi mange faktorer påvirker tilstedeværelsen

af pesticider i grundvandet. Betydningen af grundvandsmagasinets rolle i forhold til

fx nedbrydningen i overjorden kan være vanskelig at bestemme.

To tilgange kan foreslås for bedre at forstå grundvandsmagasinets rolle. Den ene er at

udnytte tilstedeværelsen af nedbrydningsprodukter i forhold til moderstof (fx ved at

beregne forholdet mellem koncentrationen af nedbrydningsprodukt og moderstof). Et

lavt forhold viser at der kun er sket minimal nedbrydning, både i overjord og

grundvand. Omvendt viser et højt forhold i sig selv kun, at der er sket nedbrydning,

enten i grundvandet eller under transporten til grundvandet, men kombineret med fx

målinger af grundvandets alder, vil det måske vise sig muligt at sandsynliggøre

nedbrydning i grundvandet.

En anden tilgang er at integrere den øvrige viden om det grundvand, hvori der

detekteres pesticider. Denne viden omfatter redoxforhold (fx i form af tilstedeværelse

og koncentration af redoxparametre som ilt eller opløst jern) og information om

tidligere eksponering til pesticider. Hvis sådan viden udnyttes, fx i statistiske analyser

ix

af grundvandsfund af pesticider, kunne det måske i nogle tilfælde føre til fund af

“nedbrydningssammenhænge”. Et eksempel på en sådan sammenhæng kunne være

manglende fund af bestemte pesticider under bestemte redoxforhold, indikerende at

en nedbrydning var sket. En sådan sammenhæng kunne efterfølgende undersøges og

måske verificeres under mere kontrollerede forhold.

x

Table of contents

1 Introduction 1

1.1 Pesticide degradation in top soil 2

1.2 Thesis delimitations 5

2 Factors affecting pesticide degradability in groundwater 7

2.1 Redox conditions 7

2.2 Microbial activity 14

2.3 Concluding remarks 17

3 Pesticide transformation processes 18

3.1 Types of transformation processes 18

3.2 Dechlorination 20

3.3 Abiotic versus biological transformation 21

3.4 Adaptation 22

3.5 Concluding remarks 25

4 Fate of selected pesticides in groundwater 26

4.1 S-Triazines 26

4.2 Phenoxy alkanoic acids 37

4.3 Nitroaromatic pesticides 40

4.4 Acetanilide pesticides 43

4.5 Phenylurea pesticides 45

4.6 Benzothiadiazone pesticides 46

4.7 Nitrile pesticides 48

4.8 Concluding remarks 49

5 Discussion and Perspectives 50

5.1 Effect of the groundwater redox environment 50

5.2 Degradability in ”real world” systems 51

5.3 Interpreting ”real world” data 51

6 References 55

xi

1 Introduction

Presence of pesticides in groundwater has been reported from all over the world,

including Europe (EEA, 1999) and the US (Barbash et al., 1999). Pesticide

contamination of groundwater gives rise to concern, since groundwater is used for

agricultural and industrial purposes and as a substantial part of the drinking water

supply in several countries. From an ecological point of view, it may be an even larger

problem, since plant and animal life in ecosystems in hydrological contact with

groundwater (like meadows, streams, swamps etc.) may be affected by pesticides in

the groundwater.

Typically only low pesticide concentrations (up to a few µg L-1) have been detected in

groundwater and the findings are often limited to more shallow secondary ground-

water. The low concentration findings may indicate that attenuation processes – most

notably degradation – in the top soil are able to minimize the pollution problem (see

below), but the findings could also be the first signs of a pesticide front, inevitably

moving towards the groundwater reservoirs. Multiple processes and factors affect the

fate of pesticides. Some factors are related to the specific pesticide, some factors are

related to the specific top soil environment to which the pesticide is applied and the

transport time through the top soil and vadose zone, and some factors are related to

the aquifer which eventually ends up being contaminated with the pesticide.

To what extent the aquifer plays a role in determining pesticide degradability is the

main topic of this thesis. It is an underlying assumption that the natural redox environ-

ment is an important factor in most groundwater processes, both biological and abiotic

processes. Since transformation of organic compounds – like pesticides - often in-

clude reduction or oxidation processes, the assumption seems reasonable, and there-

fore the approach in the following is to compare pesticide degradability with the redox

environment in which the pesticide is present. At the same time it is acknowledged

that other groundwater related factors, e.g. the groundwater pH, the presence of diffe-

rent organic and inorganic species in the aquifer and the presence and activity of

1

“competent biomass” may very well be paramount to redox in controlling pesticide

fate in aquifers.

1.1 Pesticide degradation in top soil

Several factors influence if pesticides end up in groundwater, including the amount

and type of pesticide applied in a given area, weather conditions, the hydro- and

geochemical characteristics of the soil and degradability and sorption characteristics

of the specific pesticide. Moreover, the factors may influence each other. Heavy rain

in connection to a recent application may give rise to transport of a large pesticide

amount through the top soil. Sorption as well as degradability is a function, not only

of the specific pesticide, but also of the geochemical and microbial characteristics of

the top soil.

Barbash et al. (1999) showed a good relation between the agricultural use and the

frequency of detection in shallow groundwater for five pesticides in the US (Fig. 1)

and in addition showed that the frequency of detection was lower for pesticides with

lower aerobic soil half-lives (t½).

2

(figure not available: Refer to original reference)

Fig 1. Frequencies for five pesticide detections in shallow groundwater for the 39 NAWQA

(National Water-Quality Assessment) studies in relation to agricultural use within a 1-

kilometer radius surrounding all sites samples for each study. Numbers in brackets are

networks with no pesticide detections and zero estimated agricultural use (assigned an

agricultural use value of 0.001 kg (km2)-1). a.i., active ingredient; R2, coefficient of

determination for simple linear correlations; ρ, Spearman rank correlation coefficient; t½, half

life for herbicide transformation in aerobic soil; *, correlation significant at the P<0.05 level;

**, correlation significant at the P<0.001 level; usnkluscr, study area code, Upper Snake

River Basin study area. From Barbash et al., 1999.

3

Atrazine, with an aerobic soil half-life of 146 days, was detected in more than 80% of

samples from areas with agricultural use of more than 10 kg a.i. (active ingredient)

per square kilometer, whereas cyanazine, with an aerobic half-life of 17 days, was

detected in less than 20% of samples from areas with the same use of cyanazine.

Sorption was not found to be significantly correlated to the frequency of findings of

the five pesticides (Barbash et al., 1999), probably due to a narrow range of KOC-

values, but for a larger group of pesticides with a broader range of KOC-values Kolpin

et al. (1998a) found a significant, inverse, relation between KOC and the frequency of

detections in shallow groundwater.

The presence of degradation products in groundwater also shows the importance of

degradation in top soil. Kolpin et al. (1998b) compared for five pesticides the ratio of

total product detection frequency (parent compound plus degradation products) to

parent compound detection frequency, with the top soil dissipation half-life, and

found that the faster the dissipation the higher the ratio of total product to parent

product (Fig. 2).

(figure not available: Refer to original reference)

Fig. 2. The relation between soil dissipation half-life of the parent compound and the ratio of

the total product to parent detection frequencies (TPPR). From Kolpin et al., 1998b.

4

Adams and Thurman (1991) proposed that the ratio of the groundwater content of the

atrazine metabolite deethylatrazine and atrazine could be used to assess the top soil

residence time of atrazine. The proposal was based on considerations that a high

deethylatrazine to atrazine ratio (DAR) was due to long residence time in the top soil

and thereby to an increased transformation of atrazine to deethylatrazine by soil

microorganisms, in contrast to a low DAR in situations where the residence time was

low. The DAR has been used succesfully to describe water transport to the

Mississippi river, showing that the DAR increased in base flow situations, but

decreased in surface run-off situations, where the contact time between atrazine and

the top soil were shorter (Meyer et al., 1996). Recently, the approach of using ratios

of pesticides and metabolites for evaluating infiltration rates was proposed for another

pesticide, the acetanilide pesticide metolachlor (Phillips et al., 1999).

1.2 Thesis delimitations

The presence of pesticides in groundwater can be viewed as a health problem, as an

ecotoxicological problem and as a regulatory problem. The EEC threshold limit for

pesticides in drinking water is 0.1 µg L-1 (EEC, 1980). This limit is also used for

groundwater for regulatory purposes, but is not chosen from a toxicological view-

point. In other parts of the world, higher threshold limits for pesticides in groundwater

exist. Toxicological and ecotoxicological impacts of pesticides in groundwater are not

presented or discussed in this thesis.

The sources of pesticide pollution of groundwater are still a matter of dispute. It is not

known if correct agricultural practice use of pesticides or pesticide spills and

accidents are the main reason for the pollution. Probably both contribute. In the

context of this thesis, pesticide pollutions will be viewed as infrequent non-point

source pollutions of low concentrations. Thereby pesticide pollutions are different

from point source pollutions in several ways. The sources are unspecified, the load to

the aquifer over time is not constant and may occur in irregular pulses, determined by

pesticide application and weather events, and the concentrations are in most situations

5

probably too low to alter the natural conditions of the aquifers (including redox

conditions).

Other xenobiotic pollutions may be characterized in the same way. Small amounts of

waste water may leak from old sewers to the groundwater, containing complex and

chemically diverse compounds like detergents and medicine residues. Such com-

pounds may be subject to the same aquifer processes that are described in the

following chapters.

6

2 Factors affecting pesticide degradability in groundwater

The factors affecting pesticide transformation processes in groundwater are, to a great

extent, the same which are responsible for transformation processes in top soil or

other environments. The main differences lie in the special characteristics of

groundwater that may differ from other environments.

Many groundwater factors affect transformation processes, amongst these are redox

conditions, presence of microbial life and activity, temperature and pH. Higher tempe-

ratures may increase abiotic transformation process rates, whereas temperature optima

exists for biological processes. Many hydrolysis reactions may be enhanced by high

and low pH values (e.g. fast abiotic atrazine hydrolysis to hydroxyatrazine). pH may

affect both sorption behaviour and microbial availability of some pesticides, e.g.

acidic pesticides such as the phenoxy alkanoic acids, since the chemical distribution

between neutral and anionic forms is pH controlled. Groundwater redox conditions

and microbial activity will be discussed in the following sections.

2.1 Redox conditions

The term ”redox conditions” refers to the microbial terminal electron accepting pro-

cesses taking place within the aquifer. If oxygen is present aerobic conditions will

dominate and microbial metabolism takes place with oxygen as the terminal electron

acceptor. If oxygen is depleted other electron acceptors will be used and anaerobic

terminal electron accepting processes (TEAPs) occur (e.g. Stumm and Morgan, 1996).

Nitrate, oxides and hydroxides of manganese(IV) and iron(III), sulfate and carbon

dioxide can be used as electron acceptors in order to gain energy for microbial main-

tenance and growth.

Discrete groundwater chemistry zones, indicative of different redox environments, are

often found along flow lines in aquifers, going from more oxidized to more reduced

7

conditions (e.g. Lovley and Chapelle, 1995; Appelo and Postma, 1993). Figure 3

shows a conceptual example, where the groundwater composition changes from

oxidized to reduced conditions over depth, assuming a downward vertical flow

gradient.

(figure not available: Refer to original reference)

Fig. 3. The sequence of reduction processes as reflected by groundwater composition.

Adapted from Appelo and Postma, 1993, p 258.

The zonation can be explained by energy yield considerations. Microorganisms

capable of using the more oxidized electron acceptors will have an advantage over

other microorganisms, because they gain more energy from the TEAP. Therefore the

most energy yielding electron acceptor will be depleted prior to the use of the next

best electron acceptor. In practice, this microbial competition can take place via

physiological controls like competitive exclusion processes. For example, iron

reducing bacteria are able to maintain the concentration of hydrogen and low

molecular acids at concentrations too low for sulfate reducing or methane producing

bacteria to exploit (Chapelle and Lovley, 1992). In these considerations it is assumed

that the bacteria responsible for the anaerobic TEAPs are only able to metabolize

8

simple organic compounds which are produced from more complicated structures by

fermentating bacteria (Lovley and Chapelle, 1995; Chapelle, 1993).

Abiotic processes may help in creating seemingly discrete redox zones. Manganese

oxides may be reduced by Fe2+ produced by iron reduction, leading to removal of dis-

solved Fe2+ and abiotic production of dissolved Mn2+, which indicate a solely

manganese reducing environment. In addition, Fe2+ may be precipitated by sulfides if

sulfate reduction also occurs, and methane produced in a methanogenic environment

may be used by sulfate reducing bacteria. In such cases water chemistry may indicate

that only sulfate reduction is occurring.

It has been proposed (Postma and Jakobsen, 1996) that a partial equilibrium model

gives a better and more precise explanation for redox zonation. Often the initial fer-

mentation is the rate limiting step, and by considering the TEAP to be fast and in ther-

modynamical equilibrium it is possible to predict which TEAPs are possible under

which conditions, using a number of assumptions. Using the partial equilibrium

approach, interactions between iron and sulfate reduction processes in a number of

aquifers were explained (Postma and Jakobsen, 1996; Jakobsen and Postma, 1999).

Using changes in pH, sulfate concentrations and stability of the iron oxides in the

aquifers it was possible to explain why, in certain environments, sulfate reduction

dominated over iron reduction, even though iron oxides were still present. Overlaps

between iron and sulfate reduction zones are often found (e.g. Jakobsen and Postma,

1999; Ludvigsen et al., 1998; Pedersen and Albrechtsen, 2000, I), and may be ex-

plained by the partial equilibrium model.

A large number of investigations have focused on redox environments in connection

with point source contaminations. A recent review is Christensen et al. (2000). In

such situations a high amount of organic matter, either in mixtures of a relatively few

compounds – e.g. contaminations with chlorinated solvents or refined mixtures of

hydrocarbons like gasoline – or in more complex mixtures – e.g. landfill leachates or

contaminations with crude oil – is introduced to the groundwater. This changes the

natural redox environment in a more reduced direction and redox zonation is often

seen. Closest to the contaminant source methane producing conditions are found, and

further downgradient sulfate reduction, iron and manganese reduction, denitrification

9

and aerobic conditions can be found, provided that the individual electron acceptors

are present naturally in the aquifer.

The number of investigations concerning the redox environment of pristine, uncon-

taminated aquifers are more sparse and often focus on the processes involved in

specific redox processes. For example, Pedersen et al. (1991), Postma et al. (1991),

Francis et al. (1989) and Trudell et al. (1986) focused on denitrification both in

pristine aquifers and aquifers contaminated by nitrate from agricultural use, and

Lovley et al. (1990) focused on iron reduction processes in deeply buried pristine

aquifers. The interactions between different redox processes have also been described,

e.g. interactions and competitions between iron and sulfate reduction in several

pristine aquifers (Brown et al., 1999; Jakobsen and Postma, 1999; Postma and Jakob-

sen, 1996; Chapelle and Lovley, 1992). Pedersen and Albrechtsen (2000, I) investi-

gated seven pristine and shallow Danish aquifers, and found highly varying redox

conditions, both between aquifers and within the individual aquifers. Figure 4 shows

water redox chemistry profiles from the seven aquifers, which together with sediment

chemistry analysis and TEAP bioassays provided evidence for all anaerobic TEAPs

(denitrification, manganese reduction, iron reduction, sulfate reduction, methane

production) occurring in one or more of the investigated aquifers, sometimes

simultanously.

10

Fig. 4. Redox chemistry profiles from seven pristine Danish aquifers. Dotted lines indicate

water table (mbs: meters below soil surface). Grey boxes indicate low permeable silt/clay

layers. From Pedersen and Albrechtsen, 2000a, I.

11

Several factors influence the kind of redox environment found in a given aquifer. Old

groundwater is more influenced by aquifer processes and tends to be more reduced

than young groundwater, since oxygen and nitrate may more likely be depleted. The

age is controlled by hydrogeological characteristics of the aquifer, e.g. the hydraulic

conductivities, presence of low or high permeable aquifer material layers and the

distance to water divides, which together with the initial water input to the aquifer

(e.g. from rain or river infiltration) influence flow patterns, and thereby the ground-

water age.

The organic load of the aquifers – either sediment bound or present in the water –

serve as an electron donor in the redox processes. All other factors being equal, this

would lead to more reduced environments in aquifers containing high amounts of

organic matter. In two aerobic aquifers and a denitrifying and manganese and iron

reducing aquifer TOC (total organic carbon) was generally in the range of 50 – 200

µgC (g dw - dry weight)-1 and NVOC (non-volatile organic carbon) was in the range

of 1 to 5 mgC L-1 (Fig. 5). In more strongly reduced aquifers (iron and sulfate

reducing) TOC was in the range of 200 to 2600 µgC (g dw)-1 and NVOC ranged from

2 to 63 mgC L-1. These ranges corresponded well with ranges found for other

anaerobic shallow and pristine aquifers. Arildskov et al. (2000, IV) found DOC

(dissolved organic carbon) values of 8-10 mgC L-1 in a shallow iron and sulfate

reducing and methane producing aquifer, and Jakobsen and Postma (1999) found

DOC values in the range of 0.5 to 4 mM, corresponding to 6 to 48 mgC L-1, under

similar redox conditions.

12

Fig. 5. Comparison between redox environments from seven pristine aquifers and NVOC

(top) and TOC (bottom). Samples harbouring more than one redox process, have been

ascribed an “average” redox environment for simplicity. From Pedersen and Albrechtsen,

2000a, I.

Both TOC and NVOC/DOC are crude measures of what might be available for

metabolic activity. Perhaps only a few percent of the naturally occurring carbon is

bioavailable (van der Kooij et al., 1982). None the less, it could be speculated that the

total organic content in general would give a first impression of the redox

environment, at least for shallow aquifers with relatively simple flow patterns.

13

2.2 Microbial activity

Microbial activity is a relevant measure when assessing the potential for biological

transformation processes. Microbial activity in groundwater can be highly variable,

even under the same redox conditions, and is influenced by several factors, including

the nutrient supply, pH and temperature.

Several methological approaches can be applied when studying microbial activity, and

they have different advantages and limitations. A recent review of the different

approaches, together with their potentials and limitations, can be found in Kieft and

Phelps (1997). The specific approach is important to take into acount when assessing

the specific results, and will therefore be adressed briefly in the following.

Enumeration of total number of microorganisms, of viable microorganisms or of

cellular compounds directly involved with metabolic activity (e.g. ATP) is often used

to indicate presence of microbes, and have been of great value in studies showing the

general presence of bacteria in aquifer environments (e.g. Boone et al., 1995).

Characterization of microbial populations e.g. in terms of physiological or

phylogenetic analysis may show the diversity amongst microorganisms (e.g. Balkwill

and Boone, 1997). However, in terms of microbial activity and significance, the mere

number or identity of the microorganisms says little, even when measuring metabolic

activity parameters.

Measuring changes in the concentration of relevant compounds over time is another

approach to assess microbial activity, because activity rate estimates and constraints

can be assessed. The compounds in question can include redox process species (e.g.

oxygen or Fe2+) or organic compounds. Transformation of specific xenobiotics or

more general model compounds can be followed over time, both in laboratory or field

incubation experiments, using more or less advanced approaches (Kieft and Phelps,

1997; Christensen et al., 2000). Changes in the ”real world” concentration of geo-

chemical parameters or of contaminants over time or distance may provide knowledge

of average metabolism rates, or provide constraints on the minimum or maximum

possible rates (Chapelle et al., 1995; Phelps et al., 1994; Chapelle and Lovley, 1992;

14

Chapelle and Lovley, 1990). For example the presence of oxygen at high depths in an

aquifer may be used to calculate a maximum rate of aerobic metabolism (Kieft and

Phelps, 1997) and the presence of a xenobiotic compound in ”old” groundwater may

indicate recalcitrance of the compound (e.g. Eades, 1992).

TEAP-bioassays are useful to explore whether anaerobic electron accepting processes

occur, because several of the water soluble and sediment bound chemical redox

indicators (e.g. nitrate, Fe2+, Mn2+, sulfide, FeS2, methan) are influenced by other

processes, like variation in input, transport from other parts of the aquifer, or

precipitation. For example, dissolved Fe2+ may indicate iron reduction, but is stable

under reduced and non-sulfidogenic conditions and therefore may be found under

non-iron reducing conditions. In fact, Pedersen and Albrechtsen (2000, I) observed

active iron reduction in fewer aquifers by using TEAP-bioassay than was expected

from the presence of reduced iron in the groundwater. TEAP-bioassays can be used to

assess occurrence and rates of TEAPs, and to detect both dominant TEAPs and less

dominant TEAPs occurring simultanously in both pristine and contaminated aquifers

(Pedersen and Albrechtsen, 2000, I; Cozzarelli et al., 1999; Ludvigsen et al., 1998).

TEAP-bioassays can also be used to evaluate limiting factors for TEAPs (e.g.

presence of electron donors or acceptors) (e.g. Albrechtsen et al., 1995). TEAP-

bioassays are discussed in more detail in Christensen et al. (2000).

When the potential for degradation of xenobiotic compounds (like pesticides) is in

question, a common approach is to incubate groundwater material samples with the

xenobiotic compound and follow the eventual concentration decrease over time. This

can be done both in laboratory incubations and field injection experiments. 14C-

labelled compounds give the additional possibility of also monitoring eventual

mineralization in terms of 14CO2 production. A number of investigations using both

non-labelled and labelled pesticides for transformation studies will be described in the

following chapters.

The general microbial activity, in terms of e.g. the turn-over potential rates of

generally occurring organic matter in groundwater can be assessed using model

compounds, e.g. readily degradable compounds like glucose and acetate (e.g. Kieft et

al., 1995; Phelps et al., 1994; Chapelle and Lovley, 1990; Phelps et al., 1989) or more

15

recalcitrant compounds like phenol or aniline (e.g. Swindoll et al., 1988a, b).

Pedersen and Albrechtsen (2000, II) used ring-labelled 14C-benzoic acid (benzoate)

for this purpose, because benzoic acid is an intermediate in the transformation of both

naturally occurring compounds like lignin (Young and Fraser, 1987) and several

xenobiotics (Alvarez and Vogel, 1995; Lovley and Lonergan, 1990; Kuhn et al.,

1988). It was verified that microbial activity was in fact present in the pristine aquifers

investigated, in terms of turn-over of benzoic acid. More important, the substantially

slower benzoic acid turn-over in the most reduced aquifers correlated just as well to

the generally higher contents of natural organic matter in these aquifers than to redox

conditions (Figs. 6, 7), and presence of lag phases, even at an inititial concentration of

1 µgC L-1, showed adaptation. It was therefore speculated that higher organic matter

availability in the anaerobic aquifers slowed down the adaptation to benzoic acid, and

therefore, that the turn-over rates could not be used to quantify microbial activity.

Fig. 6. 50% degradation time (DT50) of 14C-benzoic acid as a function of redox environment

for seven pristine Danish aquifers. Samples with more than one redox process have been

ascribed “average” redox status for simplicity. From Pedersen and Albrechtsen, 2000b, II.

16

Fig. 7. 50% degradation time (DT50) of 14C-benzoic acid as a function of NVOC (mgC L-1)

for seven pristine Danish aquifers. From Pedersen and Albrechtsen, 2000b, II.

2.3 Concluding remarks

Multible methods exist for evaluating both redox conditions and microbial life and

activity of aquifers. The different investigation methods have advantages and

limitations, and the results are often influenced by the methods used. It may therefore

generally be necessary to use a span of methods to supplement each other.

The number of investigated aquifers are slowly increasing, revealing an increasingly

complex picture of the subsurface. This complexity must be taken into account when

fate of xenobiotics in groundwater is to be considered. An example of this is that even

though a certain pesticide is degradable under certain conditions in one aquifer (like

specific redox conditions), degradation in other aquifers under the same conditions

does not neccessarily take place.

17

3 Pesticide transformation processes

3.1 Types of transformation processes

It is convenient to roughly differentiate between oxidative, reductive and hydrolytic

transformation processes. Other processes (e.g. isomerization, molecular rearrange-

ments) as well as conjugation processes and formation of bound residues may also

contribute to pesticide transformation. For a more detailed description of the different

types of transformation processes, see Scheunert (1992).

Examples of oxidative processes are decarboxylation of the phenoxy alkanoic acids,

e.g. 2,4-D, to the corresponding phenols (Evans et al., 1971; Don et al., 1985), or N-

demethylation of phenylurea pesticides, e.g. isoproturon (Mudd et al., 1983). Reduc-

tive processes cover reduction of nitro substituents or complete elimination of the

nitro groups, e.g. in pentachloronitrobenzene (Scheunert, 1992; Macalady et al.,

1986). Hydrolytic processes include carbolylic and sulphate ester hydrolysis, and

nitrile hydrolysis, e.g. of dichlobenil to the corresponding amide (Verloop, 1972).

Transformation may be influenced by the redox conditions. 2,4,5-T was dechlorinated

under methane producing conditions in samples from the leachate contaminated

aquifer of the Norman landfill, Oklahoma (Gibson and Suflita, 1990) while sulfate

inhibited dechlorination, presumably by changing the redox environment to a sulfate

reducing environment. Several investigations (Smelt et al., 1995; Kazumi and

Capone, 1995; Ou et al., 1988) report oxidation of aldicarb under aerobic conditions

to aldicarb sulfoxide and subsequently to aldicarb sulfone (Fig. 8). Under anaerobic

conditions hydrolytic transformation of both aldicarb and the oxidized metabolites

produces the less toxic oxime and nitrile metabolites (Kazumi and Capone, 1995).

18

Fig. 8. Chemical structure of aldicarb and aldicarb metabolites under aerobic and anaerobic

conditions.

Some pesticides are transformed by conjugation processes, where a chemical reaction

with naturally occurring organic compounds produces typically larger molecules.

Acetanilide pesticides, e.g. alachlor, are subject to glutathione conjugation in higher

organisms like plants and insects. The process results in the formation of

ethanesulfonic and oxalinic acids (Field and Thurman, 1996). Acetanilide ethane-

sulfonic acids, like alachlor ESA, are often found in groundwater in the US (Kalkhoff

et al., 1998).

Bound residue formation of pesticides are reported for top soil environments and can

be defined, rather arbitrarily, as pesticide residues not extractable from soil using

solvents (Scheunert, 1992; Khan, 1982). Mechanisms leading to the formation of

bound residues include physical deposition into clay minerals or in cavities between

19

naturally occurring organic compounds, or chemical binding to organic compounds.

Formation of bound residues can occur fast. For example, one day after atrazine

application to top soil lysimeters, 22% was non-extractable (Barriuso and Koskinen,

1996). In groundwater the importance of bound residue formation is unknown. It

could be speculated that the low content of organic matter generally found in aquifers

would limit the role of bound residue formation. However, slow atrazine removal over

time in laboratory incubations with groundwater and sediment from a number of

Danish aquifer could be explained by formation of a non-extractable fraction of the

applied atrazine (Pedersen and Albrechtsen, 2000, V – see next chapter).

3.2 Dechlorination

Pesticides are chemically very diverse compounds, but chloro constituents are often

present. Dechlorination is an important transformation process for many pesticides,

since chloro constituents often make the compound more recalcitrant. The removal of

chloro constituents may often enhance degradability of the dechlorinated metabolite

(Scheunert, 1992). Dechlorination can occur via all the above mentioned processes,

but the resulting metabolites may be different. For example, DDT (Fig. 9) is reduc-

tively dechlorinated to DDD (Glass, 1972; Zoro et al., 1974), or dehydrochlorinated to

DDE (Guenzi and Beard, 1976; Wedemeyer, 1967). Atrazine is hydrolytically dechlo-

rinated to hydroxyatrazine, whereas reductive dechlorination has not been reported.

Hydrolytic dechlorination of alachlor leads to a large number of dechlorinated meta-

bolites in groundwater samples from Massachussets (Potter and Carpenter, 1995), all

chemically different from alachlor ethanesulfonic acid formed by the dechlorination

glutathione conjugation process (Field and Thurman, 1996).

20

Fig. 9. Chemical structure of DDT and different DDT metabolites.

3.3 Abiotic versus biological transformation

Differentiating between abiotic and biological pesticide transformation processes can

be relevant. Biological enzymatic activity may enhance degradation of pesticides not

subject to abiotic transformation processes, due to lowering the activation energy

thresholds of the processes (Scheunert, 1992) and are more important in terms of

complete mineralization – and thereby to complete elimination - than abiotic pro-

cesses (Alexander, 1981).

However, abiotic and biological processes can often be difficult to distinquish from

each other, even in laboratory studies. In order to differentiate between abiotic and

biological processes, controls can be sterilized by a number of methods (e.g.

amendment with microbial inhibitors or applying heat or radiation), but the chemical

and physical properties may be changed as well (Skipper et al., 1996; Wolfe and

Macalady, 1992). In addition, many processes can occur abiotically as well as

biotically. For example, the hydrolytic dechlorination process of atrazine to hydroxy-

atrazine can occur both abiotically and microbially mediated (Kaufmann and Kearney,

1970; Skipper et al., 1967), and the hydrolytic transformation of the nitrile pesticide

21

dichlobenil to the metabolite 2,6-dichlorobenzamide (BAM) can occur by both abiotic

and biological processes (Heinonen-Tanski, 1981).

3.4 Adaptation

Adaptation and lag phases may occur before microbial transformation processes

begin, while abiotic processes may start immediately. High pesticide concentrations

or preexposure to pesticides may therefore be important for biotic transformation, in

contrast to abiotic transformation processes, which are not enhanced by preexposure.

In fact, preexposure may in extreme situations limit further abiotic transformation,

e.g. if the factor responsible for the abiotic transformation process is being depleted.

Arildskov et al. (2000, IV) found a decrease over time in the capacity of an anaerobic

aquifer to reduce the pesticide DNOC, and attributed this to a slow renewal rate of

reactive >Fe(III)-Fe2+ surfaces of the aquifer material. The load of pesticide to the

aquifer in this case was probably much higher than in most ”real world” situations.

When low concentrations of pesticides are in question, the abiotic transformation

capacity of the aquifer is probably not depleted.

Microbial adaptation to pesticides have been reported in terms of lag phases followed

by faster degradation rates. Multible groundwater investigations show degradation of

phenoxy alkanoic acids after lag phases, both in laboratory (Heron and Christensen,

1992; Klint et al., 1993; Tuxen et al., 2000; Pedersen et al., 2000, III) and field

studies (Agertved et al., 1992; Broholm et al, 2000a). de Lipthay et al. (2000) docu-

mented adaptation of bacteria to phenoxy alkanoic acid degradation in an aquifer,

previously exposed to pesticides during a seven month injection experiment. Parts of

the aquifer exposed less or not previously exposed to pesticides showed a slower

degradation potential or none at all.

The concentration level may have been an important factor in the investigations

described above, since all - for analytical reasons - were performed with relatively

high initial pesticide concentrations, compared to what is normally found in

groundwater. Toräng et al. (2000) showed the importance of the initial concentration

22

level in terms of adaptation. For 2,4-D in samples from an aerobic sandy aquifer it

was necessary to use an initial concentration of 1 µg L-1 or less in order to observe an

exponential decrease of 2,4-D, whereas degradation curves at higher concentrations

clearly showed lag phases and subsequently much faster degradation rates (Fig. 10).

In samples from another sandy aquifer, lag phases were followed by fast degradation

for initial concentrations of p-nitrophenol in the range of 31 to 529 ng g-1, thereby

indicating adaption (Aelion et al., 1987). At a lower initial concentration (14 ng g-1)

no adaptation occurred (Fig. 11).

Fig. 10. Degradation of 2,4-D at different initial concentrations under aerobic conditions in

groundwater spiked with sediment fines from a sandy aquifer, Vejen, Denmark. From Toräng

et al., 2000.

23

(figure not available: Refer to original reference)

Fig. 11. Percent p-nitrophenol respired for different soils, Lula, Oklahoma. A: Lula soil

9MM2 at 14 ng g-1 (__▲__), 529 ng g-1 (--✶ --), B: Lula soil 9NN6 at 31 ng g-1 (__▲__), 452 ng

g-1 (--✶ --), C: Lula soil 9NN7 at 485 ng g-1 (__▲__). ± standard deviations (three or four

samples). From Aelion et al., 1987.

24

In summary, there is a potential for adaptation of groundwater bacteria to pesticide

degradation – at least for phenoxy alkanoic acids. Adaptation may be important when

dealing with aquifers that have been subject to a prolonged pesticide contamination

and/or contaminations with high concentrations of pesticide, e.g. from a point source.

Whether adaptation processes are important in non-point source situations, where the

contamination occurs in pulses and in most cases in low concentrations, is not known.

3.5 Concluding remarks

If a reductive process is the rate limiting step for pesticide transformation,

transformation may be faster under anaerobic conditions. The opposite may be true

when oxidative processes are rate limiting. For some pesticides, the redox conditions

of the aquifer may therefore be important.

Biological transformation processes necessitates presence of ”competent biomass”,

which is controlled by several factors, e.g. the redox environment. Even though high

microbial activity may enhance biological transformation processes, high activity may

also lead to more reduced conditions and thereby to changed transformation path-

ways.

25

4 Fate of selected pesticides in groundwater

4.1 S-Triazines

Several pesticides belong to the s-triazine group, amongst those atrazine, simazine,

propazine and cyanazine (Fig. 12). The most used and most frequently detected s-tria-

zine in groundwater is atrazine, and the focus of this section will be solely on atrazine.

In addition, atrazine is one of the most investigated pesticides, both in top soil and

groundwater environments. The number of investigations on the transformation pro-

cesses of atrazine under different conditions and in different environments give a

diverse picture of the potential fate of atrazine.

Fig. 12. Chemical structure of different s-triazine pesticides.

26

The primary transformation of atrazine is either via dealkylation or hydroxylation,

producing deethylatrazine (DEA), deisopropylatrazine (DIA) or hydroxyatrazine

(HA) (Fig. 13). Secundary transformation processes occur via additional dealkylation

or hydroxylation, producing degradation products like deethylhydroxyatrazine or de-

ethyldeisopropylatrazine, or by deamination producing ammelines and subsequently

ammelides and cyanuric acid (Erickson and Lee, 1989). Often the concentration of

degradation products in groundwater is larger than the concentration of atrazine alone

(Kolpin et al., 1998b; Pinsky et al., 1997; Liu et al., 1996).

Fig. 13. Chemical structure of atrazine and atrazine metabolites.

The transformation can occur via abiotic or biological processes (e.g. Kaufmann and

Kearney, 1970; Skipper et al., 1967). The different transformation processes are well

described in top soil environments (Kaufmann and Kearney, 1970; Cook, 1987;

Erickson and Lee, 1989), but also in wet land environments (DeLaune et al., 1997;

Chung et al., 1996) and submerged sediments (Mersie et al., 1998a, b). Both bacteria

(e.g. Mandelbaum et al., 1993; Radosevich et al., 1997; Yanze-Kontchou and Ge-

schwind, 1994) and fungi (e.g. Masaphy et al., 1996) are capable of atrazine

transformation. The triazine ring can be used as a source for nitrogen (e.g. Crawford

et al., 1998), but not as a source for energy, since the carbon atoms in the ring are

fully oxidized (Erickson and Lee, 1989). Therefore it is expected that energy utili-

sation only occurs in the dealkylation process.

27

Both atrazine and the primary degradation products bind to the organic phase of the

soil. DEA and DIA sorbs less than atrazine, whereas HA sorbs more (Lerch et al.,

1999; Roy and Krapac, 1994). Aging phenomenae of atrazine have been reported

from top soil investigations, due to formation of bound residues of atrazine, which are

not easily extractable (Capriel et al., 1985). In a soil lysimeter experiment, 22% of the

applied 14C-atrazine was not extractable one days after application, showing that the

non-extractable fraction may be formed fast (Barriuso and Koskinen, 1996).

The transformation processes of atrazine coupled with leaching through top soil have

been investigated in several types of top soils (Sorenson et al., 1995; Sorenson et al.,

1994; Sorenson et al., 1993). Generally, HA was the main metabolite in the top

10 cm, whereas DEA and to a lesser extent DIA were produced at higher depths.

Occurrence of HA at depths higher than 10 cm was attributed to HA formation rather

than transport from the top 10 cm, due to sorption of HA (e.g. Sorenson et al., 1995).

Atrazine is often considered recalcitrant in groundwater environments (Radosevich et

al., 1996; Adams and Thurman, 1991). A number of investigations supports this. Two

field injection experiments (Widmer and Spalding, 1995; Agertved et al., 1992),

conducted under aerobic conditions, showed no removal of atrazine within the

investigation period of up to 96 days. Papiernik and Spalding (1998) placed in situ

microcosms (ISMs) in an aerobic sand and gravel aquifer and induced denitrifying

conditions in the ISMs by adding ethanol and nitrate. No removal of atrazine, DEA or

DIA was observed within 45 days. Rügge et al. (1999) injected atrazine, together with

a number of other pollutants, in the iron and sulfate reducing zone of the landfill

leachate plume of the Grindsted landfill, but found no removal of atrazine, even after

1030 days. Parallel ISM and laboratory incubations showed similar constant atrazine

concentrations. Additionally, a number of laboratory investigations have shown no or

only minor degradation of atrazine in groundwater under both aerobic (Klint et al.,

1993; McMahon et al., 1992) and anaerobic conditions (e.g. Larsen et al., 2000a;

Larsen and Aamand, 2000).

In contrast, other laboratory investigations have shown a potential for microbial

atrazine degradation in groundwater environments, although with high initial atrazine

concentrations. Mirgain et al. (1995) investigated aerobic groundwater in batch

28

incubations without sediment and found degradation of atrazine (initial concentration

20 mg L-1) after a lag phase of 15 to 44 days. Vanderheyden et al. (1997) incubated

sandy aquifer material with atrazine at an initial concentration of 4.5 mg g-1

(corresponding to 10 mg L-1 at a water content of 45%). Atrazine was mineralized in

some but not all of the incubations, after up to 30 days lag phases. Redox conditions

were not specifically stated. Radosevich et al. (1993) found atrazine degradation in

seven out of 83 samples from a sandy aquifer, incubated aerobically at an initial

atrazine concentration of 0.1 mM (approximately 20 mg L-1).

A few number of field investigations reported disappearance of atrazine under

anaerobic conditions. Stuyfzand (1998) listed atrazine among a number of both

organic and inorganic compounds, which were removed to an extent of more than

70% after bank infiltration (BI) or artificial recharge (AR) under both “anoxic” and

“deep anoxic” conditions, defined as an environment with less than 0.5 mg/L of oxy-

gen and nitrate. de Jonge and Stäb (1998) compared atrazine fate in artificial recharge

systems under “slightly anaerobic” and “anaerobic” conditions (redox conditions not

specifically defined) and found more than 90% removal during the latter conditions.

Günther et al. (1993) found a correlation between the concentration of atrazine and

the redox potential of a river bank infiltration system, and concluded that atrazine was

eliminated under anaerobic conditions.

Common for the three cited investigations was that redox conditions were poorly

defined, and that a closer evaluation of the data presented indicated that other factors

than the redox conditions alone may have influenced the fate of atrazine, in particular

the variable load of atrazine to both artificial recharge plants and bank infiltration

facilities.

However, the reported findings correspond well to other laboratory studies. Samples

from five Dutch aquifers were investigated in laboratory incubations (van der Pas et

al., 1998). Atrazine was applied at a relatively low start concentration of 70 µg L-1.

pH, organic content of the sediment and the redox potential was measured. In the Bor-

gerswold, Papenvoort and Vierlingsbeek aquifers (Fig. 14 – top) pH was below 5.7

and the redox potential was above 430 mV, indicating fairly oxidized environments.

Organic content was less than 0.1%. Substantial amounts of atrazine were removed

29

from the water phase in these samples, but only after several years of incubation. In

the Borgerswold and Papenvoort samples the removal started immediately after the

incubation start and could be described reasonably well with first order kinetics (note

log scale). In the Vierlingsbeek samples, atrazine removal took place after a two year

period of no removal. HA was detected in samples from the Borgerswold aquifer.

(figure not available: Refer to original reference)

Fig. 14. Rate of transformation of atrazine in subsoils from five Dutch aquifers: Borgerswold,

Papenvoort, Vierlingsbeek, Genderen (sampled in 1988 and 1989), and Wassenaar 1. Average

percentage of atrazine measured (100% = 70 µg L-1), with standard error (triplicate). Lines:

Approximations by first order kinetics. Sterile: γ-radiation. Note the difference in time scale.

From van der Pas et al., 1998.

30

In samples from the Genderen aquifer (Fig. 14 – middle, note different time scale)

atrazine was removed much faster at both samplings rounds, as well as in sterile

incubations (γ-radiation). pH was neutral (6.8-7.2) in this aquifer and the redox

potential was lower (-60 to 260 mV). Organic content was 0.5 to 1.0%. In the Wasse-

naar aquifer atrazine was not removed, even after six years of incubation. This aquifer

had high pH values and the redox potential was between 270 and 480 mV and

contained less than 0.1% organic matter. The authors speculated that the relatively

high organic content of the Genderen aquifers may have catalysed atrazine transfor-

mation, and that the low redox potential of this aquifer may have made reductive de-

chlorination possible.

Similar results were obtained in samples from eight Danish aquifers (Pedersen et al.,

2000, III). The fate of a number of pesticides, including atrazine, was investigated in

laboratory batch incubations. The start concentration was 50 µg L-1 for each pesticide.

It was found that atrazine was removed from the water phase in samples from the

most reduced aquifers (iron and sulfate reducing, and methane producing) but not

under the aerobic or manganese or iron reducing conditions found in samples from

other aquifers (Fig. 15). In most cases the removal followed apparent first order kine-

tics. It was speculated that the redox conditions were controlling the fate of atrazine,

but also pointed out that the amount of sediment bound organic matter could be im-

portant, since the highest content of organic matter was found in the most reduced

aquifers. However, by similar incubation experiments with samples from the Tisvilde

Hegn aquifer, with iron and sulfate reducing and methane producing conditions

(Arildskov et al., 2000, IV), atrazine was only slightly removed within the 196 days

incubation period (Fig. 16 – middle), even though sediment bound organic matter was

0.1-0.2%.

31

Fig. 15. The concentration of atrazine as a function of time in samples from four Danish

aquifers: A: Grindsted (2.4 and 2.9 mbs: Aerobic. 6.9 and 7.4 mbs: Iron reducing). B:

Frankerup (Sulfate reducing). C: Drastrup (Sulfate and iron reducing). D: Nykøbing II

(Sulfate reducing). Dotted lines: Controls. From Pedersen et al., 2000, III.

32

Fig. 16. The concentration of A, B, 2,4,5-T; C, D, Atrazine, and E, F, DNOC over time in

samples from an anaerobic aquifer, Tisvilde Hegn, Denmark. B, D, E: Controls (amended

with 0.25 mg L-1 HgCl2). Depth below soil surface (mbs) of samples shown. 2,4,5-Tcp (2,4,5-

trichlorophenol) in samples from 3.0-3.5 mbs is a 2,4,5-T metabolite. Note different time

scale, E versus F. From Arildsskov et al., 2000, IV.

33

No lag phases were observed in the investigations, which together with lack of detec-

ting degradation products (DEA, DIA, HA) lead to the hypothesis that slow sorption

and/or formation of bound residues of either atrazine or degradation products of atra-

zine were the controlling process for the removal of atrazine. However, this hypo-

thesis was weakened by the behaviour of atrazine in mercury-amended controls where

the atrazine removal was markedly slowed down (Pedersen et al., 2000, III).

Moreover, other degradation products than the three analysed for could have been

formed, e.g. by reductive dechlorination, and the possibility of mineralization could

not be ruled out.

Therefore, additional investigations, using sediment and groundwater from the sulfate

reducing Nykøbing II aquifer, focused on specifying the fate of atrazine (Pedersen and

Albrechtsen, 2000, V). Chloroform amendment and heat-sterilization were used to

sterilize incubates parallel to mercury. Results showed a removal of atrazine in the

chloroform and autoclaved controls at the same rate as in the live incubates, whereas

mercury in this investigation also slowed the removal process down (Table 1).

Thereby it was confirmed that an abiotic process was responsible, and that mercury

apparently influenced this process by some unknown mechanism, perhaps by sorption

to the same sites as atrazine.

34

Table 1. First order rate constants (10-3 day-1) for atrazine (measured by HPLC) and atrazine

and dissolved metabolites (measured as 14C-activity) in samples from the sulfate reducing

Nykøbing II aquifer. Values in brackets are standard deviations (N=6 or 2). From Pedersen

and Albrechtsen, 2000c, V.

k (10-3 day-1)

HPLC

k (10-3 day-1) 14C-activity

Biologically active bioassays

10 µg L-1 7.8 10.4

12 µg L-1 8.2 8.6

20 µg L-1 5.8 5.5

30 µg L-1 6.4 5.2

65 µg L-1 8.3 5.0

100 µg L-1 8.9 7.6

Average (N=6) 7.6 [1.1] 7.1 [2.2]

Control bioassays

Autoclaved 3 x (121°C, 20 min) (N=2) 8.7 [2.5] 6.0 [0.8]

Chloroform (8.6 g L-1) (N=2) 6.9 [1.2] 5.2 [1.3]

HgCl2 (158 mg L-1) (N=2) 2.3 [0.2] 4.3 [1.3]

HgCl2 (15.8 mg L-1) (N=2) 4.8 [1.1] 4.6 [1.0]

Azide (2 g L-1) (N=2) 165.0 [1.2] 3.5 [1.5]

Ring-labelled 14C-atrazine was used in the investigation in order to specify the fate of

atrazine in more detail. 14CO2 was not produced, ruling out a mineralization process,

and the amount of soluble 14C-activity corresponded to the concentration of atrazine,

showing that soluble degradation products were not formed (Table 1). Approximately

20% of the applied atrazine was non-extractable in samples terminated at day 63 and

day 187 (Fig. 17), showing that formation of non-extractable residues could also take

place in low organic carbon sediment and not only in top soils. Therefore, the

hypothesis was supported that slow and irreversible sorption and an eventual

formation of bound residues controlled atrazine removal. However, the role of both

35

the redox conditions and the organic content of the aquifers was not definitely

established.

Fig. 17. Recovery of 14C-activity (14C-atrazine) in day 63 and day 187 samples from the

Nykøbing II aquifer. Removed: Removed at samplings during incubation. Remaining:

Remaining in water phase. CaCl2: Extracted by 0.01 M CaCl2. I-IV: Methanol/water (*) or

methanol/formic acid (°), respectively. Residual: 14C-activity measured by sediment

combustion after extraction. Extraction IV for day 187 samples lasted nine days. Extraction

IV for day 63 samples with methanol/water was with methanol/formic acid. Bars indicate one

standard deviation of the residual 14C-activity (duplicates). From Pedersen and Albrechtsen,

2000c, V.

In summary, although atrazine may be the most investigated pesticide in terms of

environmental fate, the reports on its behaviour under groundwater conditions are

often contradicting, showing that the role of the different processes, which may be

important under different aquifer conditions, is not well understood. Groundwater

bacteria are able to transform and mineralize atrazine, at least at high concentrations

and under aerobic conditions. In reduced and/or high-organic containing aquifers

there may be a potential for atrazine removal due to slow sorption and formation of a

non-extractable fraction.

36

4.2 Phenoxy alkanoic acids

In top soil phenoxy alkanoic acids (Fig. 18) are readily degraded (Smith, 1989).

Degradation has also been reported from wet land areas (Larsen et al., 2000b) and

marine and estuarine sediments (Boyle et al., 1999). The primary transformation

pathway under aerobic conditions is by biological decarboxylation to the

corresponding chlorophenols, followed by catechol formation which destabilizes the

ring and eventually leads to ring cleavage (e.g. Evans et al., 1971; Don et al. 1985).

Dechlorination can also occur either prior to or after decarboxylation (Evans et al.,

1971). Due to low sorption phenoxy alkanoic acids are readily transported through the

top soil and the unsaturated soil into the groundwater.

Fig. 18. Chemical structure of different phenoxy alcanoic acid pesticides.

37

Several investigations have shown the potential for aerobic degradation in

groundwater environments of the phenoxy alkanoic acids MCPP (Agertved et al.,

1992; Heron and Christensen, 1992; Klint et al., 1993; Larsen et al., 2000a; Larsen

and Aamand, 2000; Tuxen et al., 2000; Broholm et al., 2000a), dichlorprop (Tuxen et

al., 2000; Broholm et al., 2000a), 2,4-D (Tuxen et al., 2000; Kuhlmann et al., 1995),

MCPA (Kuhlmann et al., 1995) and 2,4,5-T (Kuhlmann et al., 1995). In contrast,

Pedersen et al. (2000, III) found no degradation (within 365-371 days) of MCPP and

dichlorprop in laboratory batch incubations with groundwater and sediment material

from the aerobic Grindsted and Bromme aquifers (not shown), and degradation of

MCPA, 2,4-D and 2,4,5-T in samples from the Grindsted aquifer only (Fig. 19).

Similarily, MCPP, 2,4-D and dichlorprop were not degraded in some laboratory batch

incubations with material from an aerobic sandy aquifer (de Lipthay et al., 2000), and

MCPP was not degraded within a period of 200 days in samples from a chalk aquifer

(Johnson et al., 2000).

38

Fig. 19. The concentration of phenoxy alcanoic acid pesticides as a function of time for

selected samples from two Danish aquifers, Grindsted (aerobic) and Frankerup (sulfate

reducing). A: MCPA, B: 2,4-D, C: 2,4,5-T, D: Dichlorprop. Dotted lines: Controls. From

Pedersen et al., 2000, III.

39

Dechlorination may be the first transformation step under anaerobic conditions, as

shown for 2,4,5-T in samples from a methanogenic aquifer (Gibson and Suflita,

1990). Decarboxylation may also be the first transformation step, as shown for 2,4,5-

T (Arildskov et al., 2000, IV) and 2,4-D (Gibson and Suflita, 1986), subsequently

followed by dechlorination. Pedersen et al. (2000, III) observed degradation of

dichlorprop and 2,4-D in one out of three samples from one (the Frankerup aquifer)

out of seven investigated anaerobic aquifers (Fig. 19), and no degradation (within

365-371 days) of MCPP, 2,4,5-T and MCPA in any of the investigated anaerobic

samples (not shown). Whether dechlorination or decarboxylation was the first step

was not evaluated. In the anaerobic Tisvilde Hegn aquifer (Arildskov et al., 2000, IV)

2,4,5-T was decarboxylated in samples taken from the iron and sulfate reducing zone,

but not in samples from the sulfate reducing zone located 0.5 m below the iron

reducing zone, and in samples from a methanogenic zone deeper in the aquifer (Fig.

16 – top). This indicated a connection between iron reduction and degradation of

2,4,5-T in this specific aquifer. In contrast, Rügge et al. (1999) observed no

degradation of MCPP in a 1030 days injection experiment in a landfill leachate plume

under iron and sulfate reducing conditions, and Kuhlmann et al. (1995) observed no

degradation of 2,4-D, 2,4,5-T and MCPA in aquifer columns operated at sulfate

reducing conditions.

In summary, the investigations indicate a widely distributed potential for phenoxy

alkanoic acid degradation in aquifers under both aerobic and anaerobic conditions, but

also that the potential is not generally present, and that other factors than the redox

environment may influence the fate of phenoxy alkanoic acids. The commonly found

lag phases show that microorganisms may adapt to phenoxy alkanoic acid degra-

dation. The higher degradation rates in such cases may not be relevant in connection

with low concentration contaminated aquifers.

4.3 Nitroaromatic pesticides

DNOC and dinoseb (Fig. 20) both contain nitro substituents, and therefore are expec-

ted to be subject to reduction from nitro to amino substituents, as well as by dealkyl-

40

ation (Stevens et al., 1991). Aerobic degradation of dinoseb is possible, but anaerobic

degradation is faster and may lead to complete mineralization (Stevens et al., 1991),

showing a potential for degradation in anaerobic groundwater.

Fig. 20. Chemical structure of two nitroaromatic pesticides.

In groundwater investigations DNOC was used as a model compound under both

aerobic and anaerobic conditions. In an aerobic column study, DNOC was trans-

formed after a 80 days lag phase period (Tuxen et al., 2000), and in a field injection

experiment DNOC sorbed to the sediment, apparently controlled by the groundwater

pH, but was also degraded (Broholm et al., 2000b). In laboratory batch incubations

DNOC was not degraded under aerobic conditions (Pedersen et al., 2000, III).

Under anaerobic conditions fast and apparently first order degradation of DNOC was

observed (Pedersen et al., 2000, III; Arildskov et al., 2000, IV). DNOC trans-

formation also occurred in mercury amended controls, showing the process to be

abiotic, but degradation rates were generally slower in the controls (Fig. 21,

fig. 16 - bottom).

41

Fig. 21. The concentration of DNOC as a function of time in samples from four Danish

aquifers. A: Grindsted (2.4 and 2.9 mbs: Aerobic. 6.9 and 7.4 mbs: Iron reducing). B:

Frankerup (Sulfate reducing). C: Drastrup (Sulfate and iron reducing). D: Nykøbing II

(Sulfate reducing). Dotted lines: Controls. From Pedersen et al., 2000, III.

42

It was speculated that DNOC was reduced by Fe2+ sorbed at Fe(III) (hydr)oxide

surfaces as reported for other nitroaromatic compounds (Hofstetter et al., 1999; Rügge

et al., 1998; Klausen et al., 1995), and that mercury desorbed Fe2+, thereby inhibiting

the process. However, Fe2+ was probably not the only reductant, since DNOC was

also degraded in samples with less Fe2+ present at the sediment (Arildskov et al.,

2000, IV). In these sediments other reductants, like hydrogen or reduced organic

matter, could be responsible for the transformation of DNOC. In a parallel anaerobic

field injection experiment (Arildskov et al., 2000, IV) DNOC was also transformed

under iron and sulfate reducing conditions, but with much faster rates. It was

speculated that the sediment to groundwater ratios might influence the transformation

rates, in good agreement with the >Fe(III)/Fe2+ surface process being responsible. By

normalizing first order degradation rates with sediment/water ratios field rates were

comparable with laboratory rates. It was not attempted to identify degradation

products of DNOC.

In conclusion, due to the nitro substituents, nitroaromatic pesticides seem to be readily

degradable in anaerobic groundwater environments, and apparently also in certain

cases under aerobic conditions. Whether complete mineralization occurs or whether

recalcitrant metabolites are formed is not known, but the findings show the relevance

of analyzing for reduced metabolites of nitroaromatic pesticides in groundwater.

4.4 Acetanilide pesticides

Alachlor, metolachlor and acetochlor (Fig. 22) are degradable in top soil (Stolpe and

Shea, 1995; Pothuluri et al., 1990), in aquatic environments (Graham et al., 1999) and

to some extent also in groundwater. Radosevich et al. (1993) saw a 17-57% decrease

in alachlor concentrations after 138 days in four out of 81 aquifer sediment samples

incubated aerobically. Cavalier et al. (1991) incubated groundwater samples with

alachlor and metolachlor and saw degradation after lag phases of eight months or

more. Oxygen content of the groundwater was not reported, but small contents of

nitrate (0.44-0.88 mg/L) indicated oxidized conditions. Novick et al. (1986) saw

43

transformation but not mineralization of alachlor and another acetanilide pesticide,

propachlor, in samples from an aerobic aquifer (after 47 days of incubation).

Fig. 22. Chemical structure of different acetanilide pesticides.

Degradation products were more stable than the mother compound under groundwater

conditions (Clay et al., 1997; Pothuluri et al., 1990). Potter and Carpenter (1995)

detected 20 possible degradation products from alachlor in groundwater taken from

wells installed in corn fields with the last application of alachlor three years

previously. The redox conditions of the groundwater were not reported. 30 months

later they analysed samples from the same wells, finding qualitively the same com-

pounds and indicating recalcitrance of the degradation products. Especially the sul-

fonic and oxanilic acid metabolites contribute to the total amount of the acetanilide

44

pesticides, detected from 3 to 45 times more frequently than the parent compounds

(Kalkhoff et al., 1998) and in concentrations substantially higher (Phillips et al., 1999;

Kolpin et al., 1996). Pothuluri et al. (1990) investigated degradation of alachlor in soil

from a surface to groundwater profile (surface soil, vadose sediments and

groundwater sediments) of an aerobic aquifer, and saw markedly slower trans-

formation rates going from the top of the profile to the bottom. Samples were also

incubated anaerobically (purging incubation bottles with a N2/H2/CO2 gas mixture),

and significantly slower rates were obtained under these conditions.

Phillips et al. (1999) proposed the ratio of metolachlor ethanesulfonic acid (ESA) to

metolachlor as a tool to assess infiltration rates of groundwater. Assuming that

degradation to metolachlor ESA occurs only in top soil, a low ratio would indicate

fast transport, whereas a high ratio would indicate a longer contact time to the top soil.

Groundwater degradation under both aerobic and anaerobic conditions, however,

would give large ratios, even though transport through top soil was fast.

4.5 Phenylurea pesticides

Isoproturon (Fig. 23) is readily degradable in top soils (Pieuchot et al., 1996; Issa and

Wood, 1999), and mineralization of isoproturon occurred in wet land sediments under

aerobic but not anaerobic conditions (Larsen et al., 2000b). A top soil study showed

that degradation of other phenylurea pesticides (diuron, linuron, monuron and

metoxuron) was slower than degradation of isoproturon (Cox et al., 1996), but

Vroumsia et al. (1996) evaluated the ability of soil fungi to degrade different

phenylurea pesticides, and found several strains that were able to degrade e.g. diuron

faster than isoproturon.

Fig. 23. Chemical structure of isoproturon.

45

Johnson et al. (1998) showed a potential for transformation of isoproturon under

aerobic conditions in laboratory batch incubations with groundwater and sediment

material from a chalk aquifer, with monodesmethyl-isoproturon as the primary

metabolite (Johnson et al., 2000). Mineralization did not occur. Likewise minera-

lization did not occur in aerobic or anaerobic laboratory batch incubations with

material from a sandy aquifer (Larsen et al., 2000a). No transformation of isoproturon

occurred in laboratory batch incubations with groundwater and sediment material

from eight aerobic and anaerobic aquifers (Pedersen et al., 2000, III), in a column

study with aerobic groundwater material (Tuxen et al., 2000), or in an aerobic field

injection experiment (Broholm et al., 2000a). In conclusion, the potential for

isoproturon degradation under both aerobic and anaerobic groundwater conditions

seems to be limited, at least in sandy aquifers.

4.6 Benzothiadiazone pesticides

Bentazone (Fig. 24) was degradable under aerobic conditions in top soils, both by

photolysis and microbial hydrolysis processes, whereas degradation ceased under

anaerobic conditions (Huber and Otto, 1995). Bentazone is readily transported to the

underlying sediments and groundwater due to limited sorption (Romero et al., 1996).

Fig. 24. Chemical structure of bentazone.

Bentazone was slowly degradable in samples from a number of aerobic aquifers from

the Netherlands (van der Pas et al., 1998). In the Borgerswold aquifer (Fig. 25 – top)

46

the bentazone concentration decreased to 74% after 1.6 years, but after 3.9 years only

3 % remained. In other aerobic aquifers (Papenvoort and Vierlingsbeek) first order

half-lifes were 1.5 and 2.5 years, respectively. In samples from the anaerobic Gen-

deren and Wassenaar aquifers (Fig 25 – middle and bottom) the concentration of ben-

tazone decreased only insignificantly within 5.3 years. However, when some of the

Genderen incubates were contaminated with air (open circles), bentazone transfor-

mation increased substantially. Other laboratory investigations, with durations of one

year or less, showed no transformation of bentazone under both aerobic (Broholm et

al., 2000a; Tuxen et al., 2000) and anaerobic conditions (Pedersen et al., 2000, III).

(figure not available: Refer to original reference)

Fig. 25. Rate of transformation of bentazone in subsoils from five Dutch aquifers:

Borgerswold, Papenvoort, Vierlingsbeek, Genderen (sampled in 1988 and 1989), and

Wassenaar 2. Average percentage of bentazone measured (100% = 70 µg L-1), with standard

error (triplicate). Lines represent approximations by first order kinetics. Genderen 1988:

Percentage of bentazone in individual incubations with a redox potential ( ✕ ) lower or ( O )

higher than 0-26 V. Note the difference in time scale. From van der Pas et al., 1998.

47

In conclusion, bentazone is nondegradable or very slowly degradable in groundwater.

The investigations by van der Pas et al. (1998) indicated a redox impact on bentazone

degradability, as also seen in top soil, but also pointed out that pH was high in the

anaerobic aquifers and lower in the aerobic aquifers, indicating pH to be a possible

controlling factor as well.

4.7 Nitrile pesticides

Chlorthiamid and dichlobenil (Fig. 26) are readily degradable by both abiotic but

mainly biological processes in top soil (Heinonen-Tanski, 1981; Verloop, 1972) and

only shows limited transport to groundwater due to sorption. Chlortiamid is degraded

to dichlobenil or to 2,6-dichlorbenzamide (BAM – fig. 26), which also is the primary

degradation product of dichlobenil (Verloop, 1972; Beynon and Wright, 1972). BAM

is further degraded to 2,6-dichlorobenzoic acid, but only to a limited extent (Beynon

and Wright, 1972). BAM is much less sorbing than the parent compounds and is

readily transported through the top soil to underlying sediments.

Fig. 26. Chemical structure of two nitrile pesticides and nitrile pesticide metabolite 2,6-

dichlorobenzamide (BAM).

Due to the sorption behaviour of chlorthiamid and dichlobenil only few reports on

detections in groundwater exist. Eades (1992) reported on the presence of dichlobenil

in groundwater three years after a heavy rain event had washed a recently applied

48

amount of Prefix G (containing dichlobenil as the active ingredient) through the

unsaturated zone to the groundwater. Neither the redox conditions nor the presence of

BAM at the site were reported. Pedersen et al. (2000, III) found persistence of both

dichlobenil and BAM over a period of 365-371 days in laboratory incubations with

both aerobic and anaerobic sediment from different aquifers, and persistence of BAM

under aerobic conditions was also found in other field and column studies (Tuxen et

al., 2000; Broholm et al., 2000a).

In summary, nitrile pesticides seem to be recalcitrant in groundwater environments,

regardless of redox conditions. BAM is often found in Danish groundwater (GEUS,

1999), but probably due to dichlobenil transformation to BAM in top soil, followed by

transport of BAM to the groundwater, rather than production of BAM in the

groundwater.

4.8 Concluding remarks

In general, the investigations showed that certain pesticides were degradable in

groundwater environment and other pesticides seemed to be recalcitrant in

groundwater, at least under the variety of conditions investigated. For the degradable

pesticides certain conditions were needed in order for degradation to occur.

The total number of pesticide fate investigations in groundwater are sparse, especially

given the large number of pesticides used and the high number of different pesticides

found in groundwater. The phenoxy alkanoic acids and the s-triazine pesticide

atrazine, which are amongst the pesticides most investigated, show some general

trends of behavior, but also seemingly ”conflicting” behaviour in different aquifers

and aquifer environments. Even for these ”well known” pesticides there seems to be a

need for further investigations.

Other relevant but less investigated pesticide compounds – like the phenylurea and

acetanilide pesticides – may show the same ”conflicting” behaviour if the number of

investigations are expanded.

49

5 Discussion and Perspectives

5.1 Effect of the groundwater redox environment

One of the objectives of this thesis was to investigate the role of the groundwater

redox environments in controlling pesticide fate. The redox environment seemed to

influence both abiotic and biological transformation processes for some of the

pesticides. Reductive processes occurred in terms of nitro group reduction of

nitroaromatic pesticides, like DNOC and dinoseb, and reductive dechlorination of

DDT and phenoxy alkanoic acids (e.g. 2,4,5-T). However, not only reductive proces-

ses were responsible for degradation of phenoxy alkanoic acids, since decarboxylation

could also be the initial transformation step of 2,4,5-T. Aldicarb and DDT were

degradable under both aerobic and anaerobic conditions, but the transformation

processes differed substantially, and lead to different metabolites.

The phenoxy alkanoic acids were degradable under both aerobic and anaerobic

conditions, although there might be a tendency towards aerobic transformation

occurring more frequently. In some investigations, however, aerobic phenoxy

alkanoic acid degradation did not occur, or degradation occurred sporadically even in

apparently similar samples. Several studies indicated adaptation to phenoxy alkanoic

acids and that a certain threshold exposure amount had to be exceeded prior to

adaptation.

A connection between atrazine removal and anaerobic conditions was observed, both

in laboratory experiments and in ”real world” investigations, but whether the redox

environment directly influenced atrazine fate was not conclusively determined.

Since pesticides are chemically very diverse, it may not be surprising that the different

pesticides were influenced differently by the redox environment and that some pesti-

cides seemed to be generally recalcitrant regardless the redox environment.

50

5.2 Degradability in ”real world” systems

Very few investigations deal with pesticide degradability in ”real world” aquifer sys-

tems, and observed degradability is almost solely based on results from controlled ex-

periments. High concentrations of pesticides (more than 25 µg L-1) and constant

exposure are common characteristics of most laboratory investigations, where it is

more likely that a pesticide contaminated aquifer would be characterized as being ex-

posed in irregular pulses – e.g. after pesticide field application combined with rain

events – with low pesticide concentrations (below 10 µg L-1). The observed laboratory

degradability is therefore potential degradability.

On the other hand, it could be speculated that non-degradability in a constant expo-

sure/high concentration system might also be transferable to non-degradability in a

low concentration/pulse exposure system. This seems likely for biological trans-

formation processes, as long as the high concentration is non-toxic to the microbial

population.

It should be noted, however, that non-degradability is a function of experimental setup

as well as pesticide and aquifer characteristics, and that non-degradability within the

length of laboratory experiments (typically up to a few hundred days) is to be

compared with the groundwater residence time, which can be decades or longer. Even

though e.g 95% of the applied pesticide in a laboratory experiment is still present after

an investigation period of 100 days, a removal of 5% would correspond to significant

”real world” transformation within e.g. 10 years.

5.3 Interpreting ”real world” data

In Figure 1 (Barbash et al., 1999) the agricultural use of pesticides was compared with

groundwater detections. The role of the aquifer seemed to be minor, since the

presence of pesticides in aquifers to a large extent was explainable by top soil factors

like the aerobic half-life. This is not surprising, since a simple mass balance shows

that the top soil must remove almost all of the applied pesticide. Only a very small

51

fraction of the total load of pesticide actually finds its way to the groundwater, and the

potential role of the aquifers in controlling the presence of pesticides may in most

situations be hidden by the paramount role of the top soil.

None the less, it seems as if the role of the aquifer in statistical pesticide fate investi-

gations is often neglected compared to other factors. Especially the biological proces-

ses of groundwater are not considered important. Barbash et al. (1999) listed several

factors of importance in terms of pesticide detections in groundwater (Table 2).

Aquifer factors were related solely to hydrogeology and not to chemistry or biology,

and pesticide mobility and pesticide persistence referred to top soil – not to ground-

water – characteristics.

The investigations presented in the previous chapters showed at least in some cases a

potential for pesticide degradability in groundwater. It therefore could be argued that

including groundwater characteristics – e.g. water chemistry – in statistical analysis

might reveal statistical relations of interest to pesticide fate in groundwater. In an

interesting study of the occurrence of pesticides and pesticide metabolites in the

groundwater of Iowa, pesticide occurrence was compared to well depth and the

concentration of dissolved oxygen (Kolpin et al., 1997). Both parameters were used as

“rough surrogates” for groundwater age, and a significant positive relation to oxygen

content indicated that pesticides were mainly found in young groundwater. It would

have been interesting to use the knowledge of oxygen content to assess whether the

investigated aquifers were aerobic, and whether specific pesticides were mainly found

under aerobic or anaerobic conditions.

52

Table 2. Factors associated with pesticide detections in groundwater. Adapted from Barbash

et al., 1999.

Factors associated with increased likelihood of pesticide detections

Study design: - Lower analytical detection limits

- Targeting areas of higher presumed or known

vulnerability

- Targeting areas of known or suspected contamination

Pesticide Properties: - Greater pesticide mobility (lower KOC)

- Greater pesticide persistence (lower reactivity)

Agricultural Management

Practices:

- Higher pesticide use

- Increasing proximity to pesticide application areas

- Reductions in depth or frequency of tillage

Well Characteristics: - Decreasing well depth

- Dug or driven (versus drilled) wells

- Poorer integrity of surficial or annular well seals

Hydrogeologic and Edaphic

Factors:

- Unconsolidated aquifer materials (versus bedrock)

- Decreasing depth of upper surface of aquifer

- Decreasing thickness or absence of confining layers

- Higher hydraulic conductivity

- Higher soil permeability

- Increased karstification

- Increased recharge (from precipitation or irrigation)

- Younger groundwater age

Other redox relevant compounds – nitrate, Fe2+, sulfide, methane – as well as mea-

sures of other parameters (e.g. nutrients, pH and content of organic matter) and know-

ledge of previous exposure history – could be included in statistical analysis in order

to find statistical relations. Such relations might be explainable by results from more

controlled investigations, might confirm results from controlled experiments, or –

53

perhaps most important – might give leads and ideas to new controlled investigations

of possible relations.

The absolute concentration of pesticides in groundwater is often difficult to interpret

in terms of groundwater degradation since it is also a function of pesticide load,

transport, sorption and degradation in top soil. Therefore metabolite/parent compound

ratios might be of use to assess groundwater degradability. Such ratios have been

applied for triazine and acetanilide pesticides and used to assess top soil residence

time. Using metabolite/parent compound ratios might be useful to assess groundwater

degradation, if the top soil residence time could be determined by other methods, e.g.

modelling or groundwater age determination. In young groundwater a high metabo-

lite/parent compound ratio might indicate groundwater transformation, and a low

metabolite/parent compound ratio would indicate not only low transformation in the

top soil but also low transformation in the groundwater.

Difficulties in establishing real world behaviour of xenobiotics is not only related to

pesticides in groundwater. Even for “well known” compounds like the chlorinated

solvents and BTEX compounds, and for known point source pollutions with more

complex mixtures with xenobiotics (e.g. landfill leachate plumes), the complexity of

the natural environments makes an interpretation of real world data difficult. It may

be even more difficult to assess pesticide fate, due to the chemical diversity of

pesticides and the non-point source characteristics of most pesticide pollutions. The

above mentioned approaches – integrating knowledge of aquifer characteristics in

statistical analysis and using metabolites/parent compound ratios – may still be diffi-

cult to interpret, but may be a first step towards a better understanding of pesticide

degradability in groundwater.

54

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