The Influence of Soil and Contaminant Properties on the Efficiency of Physical and Chemical Soil Remediation Methods

SOFIA JONSSON

Department of Chemistry, Umeå University 901 87 Umeå

Umeå 2009

Copyright©Sofia Jonsson ISBN: 978-91-7264-763-3 Printed by Print & Media Umeå, Sweden 2009

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TABLE OF CONTENT

ABSTRACT ......................................................................................................................... I

SAMMANFATTNING (SUMMARY IN SWEDISH) ................................................................ II

LIST OF PAPERS ............................................................................................................... III

LIST OF ABBREVIATIONS AND SYMBOLS ......................................................................... IV

1. INTRODUCTION ............................................................................................................ 1

2. POLLUTANTS ADDRESSED ............................................................................................ 5

2.1 POLYCYCLIC AROMATIC HYDROCARBONS................................................................................ 5 Chemical Properties and Environmental Fate ............................................................... 5 Toxicity .......................................................................................................................... 8 Sources .......................................................................................................................... 9

2.2 POLYCHLORINATED DIBENZO-P-DIOXINS AND DIBENZOFURANS................................................... 9 Chemical Properties and Environmental Fate ............................................................... 9 Toxicity ........................................................................................................................ 11 Sources ........................................................................................................................ 12

3. REMEDIATION METHODS FOR ORGANIC .................................................................... 15

3.1 PHYSICAL TREATMENTS .................................................................................................... 16 Separation techniques ................................................................................................ 16

3.2 CHEMICAL TREATMENTS ................................................................................................... 17 Oxidation techniques .................................................................................................. 17 Reductive dehalogenation techniques ........................................................................ 20

3.3 BIOLOGICAL TREATMENTS................................................................................................. 21 3.4 THERMAL TREATMENTS .................................................................................................... 22

Incineration ................................................................................................................. 22 Thermal desorption .................................................................................................... 22 Vitrification ................................................................................................................. 23

3.5 INNOVATIVE TECHNIQUES ................................................................................................. 23 Nano-scale degradation ............................................................................................. 23 Mechanochemical degradation .................................................................................. 24

4. AVAILABILITY OF POLLUTANTS IN SOIL....................................................................... 25

4.1 BIOAVAILABILITY AND AVAILABILITY .................................................................................... 25 4.2 EFFECTS OF SOIL CHARACTERISTICS ON AVAILABILITY/ SEQUESTRATION ...................................... 26 4.3 AGED SOILS AND SPIKED SOILS ........................................................................................... 29

5. APPLICABILITY OF SOIL REMEDIATION ....................................................................... 31

5.1 REMEDIATION METHODS APPLIED TO AGED SOILS .................................................................. 32 Chemical oxidation methods [Papers I & II] ............................................................... 32 Physical remediation methods [Papers III & IV] ......................................................... 37

5.2 INFLUENCE OF SOIL CHARACTERISTICS’ AND CONTAMINANTS’ PHYSICO-CHEMICAL PROPERTIES ON

AVAILABILITY/SEQUESTRATION ................................................................................................ 43 Chemical remediation methods [Papers I & II] ........................................................... 43 Physical remediation methods [Papers III & IV] ......................................................... 49

5.3 FEASIBILITY OF THE REMEDIATION METHODS CONSIDERED ....................................................... 52 Chemical oxidation ..................................................................................................... 52 Physical treatment ...................................................................................................... 53

6. CONCLUDING REMARKS AND FUTURE ....................................................................... 57

7. ACKNOWLEDGEMENTS .............................................................................................. 61

8 REFERENCES ................................................................................................................ 63

APPENDIX ...................................................................................................................... 77

PERSONAL THOUGHTS ON REMEDIATION IN SWEDEN. .................................................................. 77

I

ABSTRACT

A vast number of sites that have been contaminated by industrial activities have been identified worldwide. Many such sites now pose serious risks to humans and the envi-ronment. Given the large number of contaminated sites there is a great need for effi-cient, cost-effective remediation methods. Extensive research has therefore been fo-cused on the development of such methods. However, the remediation of old indus-trial sites is challenging, for several reasons.

One major problem is that organic contaminants become increasingly strongly seques-tered as they persist in the soil matrix for a long period of time. This process is often referred to as ‘aging’, and leads to decreasing availability of the contaminants, which also affects the remediation efficiency. In the work underlying this thesis, the influence of soil and contaminant properties on the efficiency of various physical and chemical soil remediation methods was investigated. The investigated contaminants were poly-cyclic aromatic hydrocarbons (PAHs), polychlorinated dibenzo-p-dioxins (PCDDs) and dibenzofurans (PCDFs).

Briefly, the results show that as the size of soil particles decreases the contaminants become more strongly sorbed to the soil’s matrix, probably due to the accompanying increases in specific surface area. This affected the efficiency of the removal of organic pollutants by both a process based on solvent washing and processes based on chemi-cal oxidation. The sorption strength is also affected by the hydrophobicity of the con-taminants. However, for a number of the investigated PAHs their chemical reactivity was found to be of greater importance for the degradation efficiency. Further, the organic content of a soil is often regarded as the most important soil parameter for adsorption of hydrophobic compounds. In these studies the effect of this parameter was found to be particularly pronounced for the oxidation of low molecular weight PAHs, but larger PAHs were strongly adsorbed even at low levels of organic matter. However, for these PAHs the degradation efficiency was positively correlated to the amount of degraded organic matter, probably due to the organic matter being oxi-dized to smaller and less hydrophobic forms. The amount of organic matter in the soil had little effect on the removal efficiency obtained by the solvent-washing process. However, it had strong influence on the performance of a subsequent, granular acti-vated carbon-based post-treatment of the washing liquid.

In conclusion, the results in this thesis show that remediation of contaminated soils is a complex process, the efficiency of which will be affected by the soil matrix as well as the properties of the contaminants present at the site. However, by acquiring thor-ough knowledge of the parameters affecting the treatability of a soil it is possible to select appropriate remediation methods, and optimize them in terms of both remedia-tion efficiency and costs for site- and contaminant-specific applications.

II

SAMMANFATTNING (SUMMARY IN SWEDISH)

Som en följd av industrialiseringen har ett stort antal förorenade områden identifierats runt om i världen. Dessa platser utgör en risk för levande organismer och omgivande miljö. Med tanke på det stora antalet förorenade områden finns ett stort behov av effektiva och ekonomiskt fördelaktiga saneringstekniker. Därför har en omfattande forskning riktats mot utvecklingen av sådana metoder. Vid sanering av gamla industri-områden ställs man dock inför ett antal utmaningar.

Med uppehållstiden ökar fastläggningen av föroreningarna i jorden. Jordar, där denna process har ägt rum, beskrivs ofta som ”åldrade jordar”. På grund av detta åldrande minskar föroreningens tillgänglighet och därmed också ofta saneringseffektiviteten. I denna avhandling har markens och föroreningarnas egenskapers inverkan på sane-ringseffektiviteten studerats för fysikaliska och kemiska marksaneringstekniker. För-oreningarna som undersöktes var; polycykliska aromatiska kolväten (PAH), polyklore-rade dibenso-p-dioxiner (PCDDs) och dibensofuraner (PCDFs).

I stora drag visar resultaten från undersökningarna, att med minskande partikelstorlek binds föroreningarna starkare till marken, vilket sannolikt beror på den ökande speci-fika ytarean. Detta påverkar både saneringseffektiviteten för en metod som bygger på lösningsmedelstvätt och flera metoder som bygger på kemisk oxidation. Bindnings-styrkan påverkas också av föroreningarnas fettlöslighet (hydrofobicitet). För ett antal av de undersökta PAH:erna har dock den kemiska reaktiviteten större betydelse för nedbrytningseffektiviteten än hydrofobiciteten. Jordars innehåll av organiska material brukar ofta betraktats som en av de viktigaste faktorerna för adsorptionen av hydro-foba föreningar. Denna effekt visade sig vara särskilt markant för oxidation av lågmo-lekylära PAHer, medan de större PAH:erna var starkt bundna redan vid låga organiska halter. Oxidationseffektiviteten av större PAHer var dock positivt korrelerad med mängden nedbrutet organiskt material. Detta beror sannolikt beror på att det organis-ka innehållet i jorden oxiderades, varvid dessa föroreningar frigjordes. Mängden orga-niskt material verkade ha liten effekt på saneringseffektiviteten som uppnåddes med lösningsmedelstvätt, men den hade större inverkan på resultatet vid en efterföljande behandling av tvättvätskan.

Sammanfattningsvis visar resultaten i denna avhandling att sanering av förorenad mark är en komplicerad process där saneringseffektiviteten påverkas av jordens matris och föroreningarnas egenskaper. Genom att skapa en övergripande förståelse för de parametrar som påverkar behandlingen av förorenad mark ökar möjligheten att välja rätt saneringstekniker. Därigenom blir det också möjligt att optimera saneringsmeto-derna med avseende på såväl effektivitet som kostnader.

III

LIST OF PAPERS

This thesis is based on the following papers, which are referred to in the text by the corresponding roman numerals

I. Sofia Jonsson, Ylva Persson, Sofia Frankki, Staffan Lundstedt, Bert van Bavel, Peter Haglund and Mats Tysklind, Comparison of Fenton’s Reagent and ozone oxidation of poly-cyclic aromatic hydrocarbons in aged contaminated soils. Journal of Soils and Sediments 6 (2006) 208-214.

ІI. Sofia Jonsson, Ylva Persson, Sofia Frankki, Staffan Lundstedt, Bert van Bavel, Peter Haglund and, Mats Tysklind, Degradation of polycyclic aromatic hydrocarbons (PAHs) in contaminated soils by Fenton's reagent: A multivariate evaluation of the importance of soil characteristics and PAH properties Journal of Hazardous Materials 149 (2007) 86-96.

ІІІ. Sofia Jonsson, Henrik Lindh, Staffan Lundstedt, Peter Haglund and Mats Tysklind. Dioxin removal from contaminated soils using ethanol washing. Submitted to Journal of Hazardous Materials

ІV. Sofia Jonsson, Henrik Lindh, Staffan Lundstedt, Peter Haglund, and Mats Tysklind. Carbon adsorption of dioxins – a possible post-treatment of ethanol used to wash contaminated soils? Manuscript

IV

LIST OF ABBREVIATIONS AND SYMBOLS

AC Activated carbon APEG Alkalimetal hydroxide/polyethylene glycol BCD Base-catalysed decomposition CP Chlorophenol DOM Dissolved organic matter GAC Granular activated carbon HMW High molecular weight LMW Low molecular weight MIFO Methodology for inventory of contaminated sites MO Microorganisms NAPL Non-aqueous liquids NOAR Number of aromatic rings NOM Natural organic matter PAC Polycyclic aromatic compound PAH Polycyclic aromatic hydrocarbon PC 1 First principal component PCA Principal component analysis PCB Polychlorinated biphenyl PCDD Polychlorinated dibenzo-p-dioxin PCDF Polychlorinated dibenzofuran PHWE Pressurised hot water extraction PHWO Pressurised hot water oxidation PLS Partial least squares projections to latent structures POM Particulate organic matter POP Persistent organic pollutant S-EPA Swedish Environmental Protection Agency SCWE Supercritical water extraction SCWO Supercritical water oxidation SOM Soil organic matter TCDD Tetrachlorodibenzo-p-dioxin TCDF Tetrachlorodibenzofuran TEF Toxicity Equivalency Factor TEQ Toxic Equivalent Quantity TOC Total organic carbon US-EPA United States Environmental Protection Agency UV Ultraviolet WHO -TEQ Toxic equivalence according to the World Health Organisation

V

SYMBOLS

5 ring Number of five-membered rings Ip Ionization potential (HOMO energy) (eV) Hvap Heat of vaporization (kcal/mol) Sw Water solubility (mmol/l) Kow Octanol-water partition coefficient Mw Mole weight (g/mol)

1

1. INTRODUCTION

During the 20th century rapid industrialization led to large-scale releases of in-dustrial chemicals worldwide. Many of the chemicals released are potentially harmful compounds that may cause adverse effects in humans and the envi-ronment. However, increasing environmental awareness has led to reductions in the release of many of these harmful substances due to prohibition of their production and use and/or changes in industrial processes. Nevertheless, the industrialization has left a legacy of contamination that we must address. In some cases the contaminants have persisted for more than a century, since many of them are resistant to degradation and often highly hydrophobic, ena-bling them to sorb strongly to particles in matrices such as soil. Many of the polluted sites around the world pose great environmental risks and are conse-quently in need of remediation. However, remediating and securing contami-nated sites is extremely costly, generally costing in the order of tens or hun-dreds of millions of dollars per site [1]. Thus, development of cost-efficient remediation methods is one of the great environmental challenges today.

In Sweden inventories of potentially contaminated sites have been conducted according to the methodology for inventory of contaminated sites (MIFO), to fulfil part of the official Swedish objective, a ‘Non-Toxic environment’ (i.e. “The environment must be free from man-made or extracted compounds and metals that represent a threat to human health or biological diversity.” [2]. Fig-ure 1 shows, as an illustration, sites in the county of Västerbotten identified as being potentially contaminated, which are clearly clustered near the coasts and rivers. The reason for this is that many of the former industries used water as an energy supplier and/or for transporting raw materials [3]. This is particularly true for former sawmills, where applied processes often included wood preser-vation. Unfortunately, a consequence of locating such sites near a coast or a river is that it facilitates rapid transportation of contaminants to the environ-ment.

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Figure 1. Identified locations of potentially contaminated soils in the county of Västerbotten.

The inventory in Sweden identified more than 80 000 sites that are estimated to be contaminated with organic and/or inorganic pollutants [4]. Approxi-mately 1500 sites are classified as risk class 1, i.e. sites posing the highest risk for humans or the environment. The high number of identified sites reflects the great need for remediation in Sweden as a result of the industrialisation. Furthermore, since Sweden is not unique in any way, it also indicates the scale of this problem in industrialized countries generally. Although we do not have a complete understanding as yet of the extent of polluted sites globally, the number of contaminated sites worldwide is likely to be vast.

Hence, there is a need for a broader approach to identify and treat contami-nated sites on a global scale [1]. The widespread concern regarding contami-nated sites, has prompted increasing efforts (reflected in increasing numbers of published papers) to develop efficient, preferably cheap remediation methods during the past 20 years. These intensive efforts have resulted in the develop-ment of a wide range of remediation techniques. However, all of these tech-niques have specific advantages and disadvantages, so none of them are suit-able for treating every contaminant, or every type of soil. Therefore, appropri-ate methods should be carefully selected for each application.

It is a well-known fact that when persistent organic contaminants enter the environment they often become sequestered in the soil. This process is gov-erned by the soil’s characteristics and the contaminants’ physico-chemical properties. The extent to which it occurs is widely debated in the literature, but

3

regardless of its extent, sequestration of contaminants increases their resistance to remediation and reduces their availability. Thus, in order to assess the appli-cability and limitations of different remediation techniques, and to select an appropriate method for a site, it is of utmost importance to elucidate how the soil and contaminant properties affect the sequestration and treatability of the soil. To date, research on the effects of sequestration has mainly focused on its impact on bioavailability or bioremediation, often from a risk assessment per-spective. Less effort has been spent on assessing its impact on the efficacy of chemical or physical remediation methods, although similar factors are likely to be involved in these processes.

The two major objectives of the work this thesis is based upon were: 1) to in-vestigate the influence of soils’ and contaminants’ properties on the treatability of polycyclic aromatic hydrocarbons (PAHs), polychlorinated dibenzo-p-dioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs) in aged soils; and 2) to investigate the feasibility of applying chemical and physical soil reme-diation techniques to various aged soils contaminated with PAHs and PCDD/Fs. These objectives were addressed in the four studies reported in Papers I-IV by:

� Applying two chemical oxidants, ozone and Fenton’s reagent, to nine

aged soils of different origins, contaminated with PAHs in varying concentrations and contamination patterns (I)

� Multivariate data evaluation of the degradation efficiency of a tech-nique involving application of relatively mild Fenton reagents to ten aged PAH-contaminated soils that were well characterised with regards to soil parameters and physico-chemical properties of the PAHs (II)

� Identifying important parameters in ethanol washing of PCDD/F-contaminated soils and applying the washing procedures to four aged soils with different soil characteristics and origins (III)

� Investigating the feasibility of using granular activated carbon materials to adsorb the PCDD/Fs in washing fluids used in ethanol washing of two soils with different characteristics (IV)

5

2. POLLUTANTS ADDRESSED

Persistent organic pollutants (POPs) are present in the environment in widely varying concentrations depending on the scale, nature and proximity of con-taminating sources [1, 5]. POPs formed as by-products of incomplete combus-tion in anthropogenic processes often spread diffusely over large areas, whereas at former industrial sites the concentrations of POPs are often consid-erably higher [5]. Important characteristics of POPs include high resistance to metabolic processes and environmental degradation, by chemical and photo-lytic processes for instance. Hence, even if the release of toxic POPs is discon-tinued, they will remain in the environment for a long time. Further, many of the most common POPs have high tendencies to bioaccumulate or biomag-nify, mainly governed by their lipophilicity, which can be described by their octanol-water partition coefficients (Kow) [6, 7].

This thesis focuses on three classes of POPs: polycyclic aromatic hydrocarbons (PAHs), polychlorinated dibenzo-p-dioxins (PCDDs), and polychlorinated dibenzofurans (PCDFs). Typical sources of these compounds, their individual characteristics and toxicity are described in more detail below.

2.1 Polycyclic aromatic hydrocarbons

Chemical Properties and Environmental Fate Unsubstituted PAHs consist of carbon and hydrogen, arranged in a varying number of fused aromatic rings in linear, angular or clustered formations (Fig-ure 2). PAHs have been intensively investigated due to their toxicity and persis-tence, and the United States Environmental Protection Agency (US-EPA) has designated 16 PAHs as “priority pollutants”. Their individual physico-chemical properties are mainly governed by their size [8]. They range from compounds that are slightly soluble in water and rather volatile (2 or 3 rings), to highly hy-drophobic and practically non-volatile compounds (≥ 4 rings).

6

Figure 2. Structures of a selection of 2-6 ring PAHs

Selected physico-chemical properties of 24 PAHs with 2-6 aromatic rings are listed in Table 1. The highly heterogenic character of the PAHs’ physico-chemical properties has a major impact on their environmental fate. Generally, as shown in Table 1, the octanol/water partition coefficient of PAHs increases with increasing molecular weight [9]. Hence, the high molecular weight (HMW) PAHs have higher affinity to soil organic matter, and are thus primar-ily transported adsorbed to particles in air and water [12, 13]. Sorption of PAHs to particles also increases their chemical stability and resistance to deg-radation [14], thus emitted PAHs can be transported long distances from their sources when sorbed to particles [15]. In fact, they are found in measurable concentrations even in remote areas like Svalbard, where local sources have been estimated to have made very limited contributions to measured concen-trations [16]. In contrast, the low molecular weight (LMW) PAHs, i.e. PAHs with two or three aromatic rings, generally have higher water solubility and higher vapour pressures [9]. They may therefore occur as both dissolved and gaseous species in the environment, and are usually considerably more mobile in the atmosphere and water than the HMW PAHs [16]. Consequently, the LMW PAHs are more susceptible to degradation processes in the environ-ment, such as microbial degradation, chemical oxidation and degradation by ultraviolet (UV) light than the HMW PAHs. In soil, these characteristics, to-gether with leaching and volatilization processes, often lead to a depletion of the LMW PAHs over time [18, 19]. Hence, naphthalenes are often found at considerably lower concentrations in contaminated soils at former industrial sites than larger PAHs [20, Papers I and II].

Naphthalene Acenaphthene Fluorene Phenanthrene

Pyrene Benzo[a]anthracene Chrysene Benzo[k]fluoranthene

Benzo[a]pyrene Dibenz[a,h]anthracene Indeno[1,2,3-cd]pyrene Benzo[ghi]perylene

Anthracene

7

Table 1. Selected properties of PAHs analysed in Paper II [9-11].

PAH NOAR 5 ring Ip-homo Hvap log Sw log Kow Mw

Naphthalene 2 0 -8.711 43.2 2.4E-01 3.4 128

2-Methylnaphthalene 2 0 -8.620 45.3 1.8E-01 3.99 142

1- Methylnaphthalene 2 0 -8.584 54.8 2.0E-01 3.97 142

Biphenylene 2 0 -8.443 48 6.2E-02 3.96 152

2-6-Dimethylnaphthalene 2 0 -8.531 1.0E-02 4.34 156

Acenaphthylene 2 1 -8.943 50.4 4.0E-02 3.95 152

Acenaphthene 2 1 -8.495 51.1 2.9E-02 4.04 154

2-3-5-Trimethylnaphthalene 2 0 170

Fluorene 2 1 -8.711 58.2 1.2E-02 4.22 166

Phenanthrene 3 0 -8.617 52.7 7.2E-03 4.5 178

Anthracene 3 0 -8.123 52.4 3.7E-04 4.46 178

1-Methylphenanthrene 3 0 -8.533 1.4E-03 5.14 192

Fluoranthene 3 1 -8.631 66.5 1.3E-03 5.2 202

Pyrene 4 0 -8.885 65.8 7.2E-04 5.06 202

Benzo(a)anthracene 4 0 -8.206 65.6 6.8E-05 5.82 228

Chrysene 4 0 -8.371 65.8 1.3E-05 5.81 228

Benzo(b)fluoranthene 4 1 -8.567 3.1E-05 6.29 252

Benzo(k)fluoranthene 4 1 -8.300 3.6E-03 6.59 252

Benzo(e)pyrene 5 0 -8.219 70.8 2.50E-05 6.92 252

Benzo(a)pyrene 5 0 -7.922 71.1 1.50E-05 6.3 252

Perylene 5 0 -7.858 71.3 1.20E-05 6.12 252

Dibenz(a, h)anthracene 5 0 -8.283 75.2 8.20E-05 7.16 278

Indeno(cd)pyrene 5 1 -8.137 276

Benzo(ghi)perylene 6 0 -8.024 75.8 2.00E-05 7 276

PAHs often co-exist with large numbers of related compounds in the envi-ronment, collectively referred to as polycyclic aromatic compounds (PACs), of which the PAHs are by definition a subgroup. The PACs are all structurally quite similar, but their physico-chemical properties may differ significantly. Carbon atoms in the benzene rings may be substituted by nitrogen, sulphur or oxygen atoms, leading to heterocyclic PACs. In addition, the PAHs may con-tain substituents, e.g. nitro-, alkyl- or carbonyl groups, on the rings. PAHs with carbonyl groups, generally referred to as oxygenated PAHs or oxy-PAHs, may be formed during degradation processes, including both natural and remedia-tion processes. Indeed, levels of some oxy-PAHs in treated soils may be com-parable to the PAH levels prior to remediation [20].

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Toxicity PAHs are often regarded as POPs since they frequently seem to be highly per-sistent when sorbed to particles, but unlike many other hydrophobic organic compounds they do not bioaccumulate in higher organisms. Instead, the PAHs are metabolically transformed into more water-soluble compounds that can be excreted by the organisms. Nevertheless, PAHs can have severe biological ef-fects, since they can be: acutely toxic; carcinogenic, and mutagenic (thus affect-ing both development and reproduction), partly because reactive epoxides are formed that can react with DNA during their metabolism [21, 22]. One of the most extensively studied PAHs is benzo(a)pyrene, since it is considered to be one of the most carcinogenic [23, 24]. Many theories have been postulated to explain why some PAHs have carcinogenic activity, but not others. However, our understanding of the reasons for the carcinogenicity of some PAHs is in-complete, and there has been little success to date in identifying relationships between PAHs’ structural features and their carcinogenicity. Nevertheless, it seems that a minimum of four rings is needed for PAHs to display carcino-genic properties [25]. For substituted PAHs the identification of structure-activity relationship is even more complex. For instance, among methyl-benz(a)anthracenes, the most potent carcinogen, is formed by substitution at position 7 and substantial activity is also observed for substitution at position 6, 8 or 12. However at positions 1, 2, 3 or 4 totally inactive compounds are formed [25], illustrating the complexity in predicting the carcinogenicity of PAHs.

Published investigations have often been limited to analyses of the 16 “US-EPA PAHs”. However, oxygenated PAHs for instance are generally signifi-cantly more mobile than PAHs, and are at least equally harmful to organisms as PAHs, since these compounds have been shown to be acutely toxic, mutagenic and/or carcinogenic [20, 26]. Oxygenated PAHs are therefore also of great potential concern. Hence, it has been suggested that oxygenated PAHs should also be monitored during the treatment of PAH-contaminated soils, since if the degradation of the PAHs is incomplete, oxygenated PAHs may be formed [20].

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Sources Polycyclic aromatic hydrocarbons (PAHs) are ubiquitous in the environment and comprise one of the most frequently occurring groups of organic con-taminants in soils and sediments.

The formation and emission of PAHs are associated with incomplete combus-tion of organic material or fossil fuels [1, 5]. They also occur naturally in coal and crude oil. Domestic heating appliances release PAHs that spread diffusely over large areas and collectively make major contributions to the total emission of PAHs in Sweden for instance [1]. However, extremely high levels of PAHs in soils are often found at industrial sites, typically including gasworks, wood-preservation, petroleum-processing/handling, metallurgic, and coke, coal-tar and creosote production sites [1].

2.2 Polychlorinated dibenzo-p-dioxins and dibenzofurans

Chemical Properties and Environmental Fate PCDDs and PCDFs are two groups of chlorinated tricyclic aromatic com-pounds with similar physico-chemical properties and characteristically planar configurations, consisting of two chlorinated benzene rings with either two (PCDD) or one (PCDF) oxygen atoms between them. The general chemical structures of PCDDs and PCDFs and the 2, 3, 7, 8 substituted tetrachlorinated (TCD) dioxin and furan are shown in Figure 3

Figure 3. General chemical molecular structure, numbering and nomenclature of PCDDs and PCDFs, and structural formulae of 2, 3, 7, 8- TCDD and 2, 3, 7, 8-TCDF.

Clx Cly

1 2

3

45

6

7 81

3

45

6

78

2

2,3,7,8 -TCDF

Clx Cly

O

OCl

Cl Cl

O

O

O

O

Cl

Cl

Cl

Cl

Cl

2,3,7,8 -TCDD

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The number of potential chlorine substituent positions (one to eight) gives rise to 210 individual PCDDs and PCDFs: 75 dioxins and 135 furans. The chlorine atom positions determine their names. Generally, PCDD/F congeners are considered to be chemically stable and persistent, but there are considerable variations in their environmental behaviour and fate, which are governed by their physico-chemical properties. Selected properties of the PCDD and PCDF congeners with four to eight chlorine atoms are shown in Table 2. Tetra- through octa-chlorinated dioxins and furans show low solubility in water, and high hydrophobicity, which respectively decreases and increases with increases in the number of chlorine atoms [27]. Congeners with a 2, 3, 7, 8 substitution pattern have the most pronounced toxic properties, partly because they are not readily metabolised, and thus can easily accumulate in fatty tissues in higher organisms [28, 29].

Table 2. Brief overview of properties of tetra- through octa-chlorinated dioxins and furans. Compounds Molecular weight

(g/mol) Log Sw Log Kow

TCDD-OCDD 320 - 456 6.16-9.64 1 6.91-8.75 1 TCDF- OCDF 304 - 440 6.91-9.6 1 6.06-8.60 1 1 Govers H., Krop H., (1996) [30].

The most common way for dioxins to enter organisms is via food intake. For instance, consumption of fish and animal products generally makes the highest contributions to overall PCDD/F exposure in humans [29]. Fatty fish, such as salmon and herring accumulate dioxins and furans particularly strongly, due to their high fat content, dioxins have been found in human milk and they have been found to be transferred to babies during pregnancy and breastfeeding [28, 31, 32]. In order to avoid this, the Swedish national food administration rec-ommends that pregnant women should limit their intake of salmon and salmon trout from the Baltic Sea to at most 2-3 times per year, and in 2001 the Euro-pean Commission banned the sale of fish from the Baltic Sea with dioxin levels exceeding those permitted by the EU to European Union countries [33].

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Toxicity The toxicity of PCDD/Fs is strongly dependent on the positions of their chlo-rine substituents. Thus, to facilitate the risk assessment and regulatory control of exposure to dioxins, the concept of Toxicity Equivalency Factors (TEFs) was developed to provide a means to equate the toxicity of dioxins and dioxin-like compounds, including some polychlorinated biphenyls (PCBs) [34]. PCDDs and PCDFs with chlorine atoms in the 2, 3, 7, 8 positions (see Figure 3), are of most concern, and 2, 3, 7, 8–tetrachlorodibenxzo-p-dioxin (TCDD) is considered to be the most toxic compound. For that reason each of the con-geners with chlorine atoms occupying all four of the 2, 3, 7 and 8 positions are among the compounds that have been assigned TEF values. TEF values of individual compounds – which were recently re-evaluated by the World Health Organisation, WHO [35] – can be used to calculate total WHO-Toxic Equiva-lence (WHO-TEQ) values, representing the total toxicity of mixtures of PCDD/Fs and dioxin-like compounds, by multiplying the congeners’ concen-trations by their respective TEF values and summing the resulting values.’

WHO-TEQs were originally intended to help assess risks posed by oral con-sumption of the compounds, and they have little toxicological relevance for contaminated soils, due to the limited bioavailability of contaminants in soil [35]. However, WHO-TEQ values are frequently used as indicators of the overall PCDD/F pollution of environmental samples. For instance, the Swed-ish Environmental Protection Agency (S-EPA) uses WHO-TEQ values per gram dry weight of soil [36] as indicators of the need for remediation, and the success of remedial treatments. For that reason, this concept is also used at various points in the discussion in this thesis and in Papers III and IV.

Dioxins have shown a number of adverse toxic effects notably after the indus-trial incident in 1976 in Seveso, Italy, where a chlorophenol production plant exploded, dioxins were spread over the neighbouring area and the workers were affected by acute symptoms, including chloro-acne. A further case of severe dioxin pollution occurred during the Vietnam War (1961-1971), when Agent Orange, which contained high concentrations of dioxins as by-products, was used as a defoliant in southern Vietnam. Today, soil and sediments near Bien Hoa City, which was near the former air base used for Agent Orange spraying missions, show elevated dioxin levels [37, 38], and Schecter et al. found extremely elevated levels of TCDD in ducks, chicken and fish, all of which are commonly eaten in the south of Vietnam [38]. This is alarming since the primary route for human exposure of dioxins is through food [29]. Schec-ter et al. also compared dioxin levels in blood samples from residents of Bien Hoa City and samples from immigrants from places where Agent Orange had not been sprayed. Up to 200-fold higher levels were found in the Hoa Bien

12

residents more than 30 years after the spraying [38]. Long-term effects of TCDD exposure include disruption of reproductive processes and carcino-genicity. For instance, maternal exposure to dioxins has been suggested to in-crease the risk for modified neonatal thyroid functions, which may result in severe mental and physical retardation [39].

Sources Polychlorinated dibenzo-p-dioxins (PCDDs) and polychlorinated dibenzofu-rans (PCDFs) are amongst the most toxic anthropogenically dispersed pollut-ants in the environment. Although dioxins have never been intentionally pro-duced for technical applications they are formed as by-products during a wide variety of processes. Contamination of soil and water by PCDD/Fs has oc-curred as a result of combustion processes, metallurgical activities, chlorine bleaching of pulp and paper and the synthesis of organochlorine agents such as pentachlorophenol (PCP) [1, 5, 40, 41]. The contamination pattern will differ depending on the source, and this can be useful for identifying sources [40]. Large- and small-scale combustion processes and other high temperature proc-esses cause a diffuse spread of PCDD/Fs [1, 5, 42, 43], whereas chlorine bleaching of pulp, wood-impregnation activities involving chlorophenol (CP) agents, and chlor-alkali production are known to have been responsible for the pollution of soil at many sites that are heavily contaminated with PCDD/Fs [1, 5]. For instance, the Swedish EPA (S-EPA) has estimated that chlorophenols have been used at 400-500 sawmill sites in Sweden [44] and that historical wood treatment alone resulted in the release of 205-250 kg TEQ [5].

As knowledge regarding dioxin formation has increased, many formerly used processes have been stopped, or modified, so PCDD/F emissions have been reduced [42]. The modifications have included changes in production proc-esses per se, the introduction of various exhaust/effluent treatments and prohi-bition of the use of some chemicals. Notably, efficient air pollution control devices have been developed to meet more stringent regulations regarding emissions during waste combustion. However, the former processes have left a toxic legacy. For instance, although the chlorine-based bleaching process has been almost completely replaced by other techniques in Sweden and most other industrialized countries [1], PCDD/Fs that were discharged into waste-waters and sludges are still present in high levels in the sediments around paper and pulp mills that used the process.

13

At sites where chlorophenols were used for wood treatment, complex con-tamination profiles, including diverse chlorinated aromatic compounds, are often found. Chlorophenols are often present at low concentrations among these compounds at former sawmill sites, due to their relatively high water solubility, which allows them to be readily transported. However, other co-contaminants that have much higher hydrophobicity, such as dioxins and fu-rans, are unlikely to be washed out from the soil, and their presence at the former sawmill sites still poses problems. These sites were historically major sinks for dioxins and furans, and they have significant relevance as secondary sources of PCDD/Fs today [1]. It has been found that secondary releases may be severely underestimated [45].

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3. REMEDIATION METHODS FOR ORGANIC POLLUTANTS

Growing numbers of remediation techniques have been developed in recent decades as many countries have recognised the problems associated with envi-ronmentally polluted areas. In this chapter, some of the available techniques for remediating soils contaminated with persistent organic pollutants (POPs) are discussed. Since the methods and applications are highly diverse not all techniques will be covered, but the techniques that are discussed here can be divided into the following groups: physical, chemical, biological and thermal treatments. The primary focus is on methods that are frequently used or are currently under development for remediating soils contaminated with organic contaminants, especially PAHs, dioxins and furans.

Remediation may be carried out either in situ (in the ground) or ex situ (after removing the contaminated material, either on-site or off-site). In addition, a technique may be used alone, as a ‘singular method approach’, or in conjunc-tion with other techniques, i.e. as part of a ‘multiple method approach’ (Figure 4)

Figure 4. Examples of possible remediation strategies. Remediation methods can be applied directly in a singular mode (■), in a dual mode (■) or in combinations of three or more remediation techniques (■).

Single method approach Physical, Chemical, Biological or Thermal Treatment

Multiple method approach

Chemical Treatment

Thermal Treatments

Biological Treatment

Separation Techniques Chemical. Biological and Thermal Treatments, Adsorbents

Biological Treatment Chemical Treatment

Single method approach Physical, Chemical, Biological or Thermal Treatment

Multiple method approach

Chemical Treatment

Thermal Treatments

Biological Treatment

Separation Techniques Chemical. Biological and Thermal Treatments, Adsorbents

Biological Treatment Chemical Treatment

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3.1 Physical treatments

Separation techniques As the term implies, separation techniques separate contaminants from the contaminated matrix by transferring them into another medium, e.g. from con-taminated soil to a liquid, or from a contaminated liquid to an adsorbent. In many cases separation techniques also reduce the volume of contaminated material.

Soil washing is an example of a commonly used separation technique. Histori-cally, soil washing has mostly been performed as an “extraction technique” using water-based solvents [46, 47]. Soil washing has been documented to work well on soils with coarse material, but less well on soils with higher silt, clay and/or organic matter contents [47, 48], because coarse soils are more permeable for the liquid, and can bind lower amounts of contaminants since they have smaller particle surface areas per unit volume [49, 50]. Hence, separating fine soil (clay and silt) from coarse soil (sand and gravel) prior to or during soil washing treatments can significantly reduce the volume of the most contami-nated materials [49-52]

Recently soil-washing processes have been modified and the use of low mo-lecular weight alcohol solvents has been shown to be a feasible method to ex-tract organic contaminants such as PAHs and dioxins from soil [53-57]. Fur-ther, by using different extracting solutions it has been shown that it is possible to extract both metals and organic pollutants such as PAHs [58, 59]. Other separation techniques that have been reported in the literature include (inter alia) the use of surfactants [60], cyclodextrins [61-62] and vegetable oils [63-65] for separating PAHs and dioxins from soil.

Pressurised hot water extraction (PHWE) or supercritical water extraction (SCWE) are extraction techniques that have to be run under certain high pressure and tem-perature conditions (water adopts a “super-critical” state when the temperature and pressure exceed critical points, at 374 °C and 221 bar, respectively). Near the critical point the properties of water change dramatically: its dielectric con-stant drops and its polarity becomes similar to that of organic solvents [47] Hence, water become very efficient in extracting even highly hydrophobic compounds such as dioxins and PAHs [66-69] and the new properties of water make HWE and SCWE an environmentally friendly remediation method.

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As previously mentioned, separation techniques only separate the contami-nants from the contaminated matrix, no degradation takes place. Therefore some kind of post-treatment of the separating media, i.e. water solutions, sol-vents or vegetable oils, is necessary. Examples of such treatments include re-generation of the solvent by distillation [70], UV-degradation [63, 64] or ad-sorption of the extracted contaminants by activated carbons [71]. Post-treatment of the residual soil may also be required for safe disposal.

3.2 Chemical treatments

Oxidation techniques Oxidation techniques involve destruction of the contaminants through reac-tion with an oxidant. Various oxidants may be used, e.g. those shown (in order of strength, from the top) in Table 3. The strongest of these reagents are hy-droxyl radicals, which are widely used in many chemical oxidation techniques.

Table 3. Comparison of oxidant strengths of different chemical specie

Chemical Species Standard oxidation

potential (Volts)

Relative strength (Chlorine= 1.0)

Hydroxyl radicals (OH•) 2.8 2.0 Sulphate radical (SO4

•) 2.5 1.8 Ozone 2.1 1.5 Sodium persulfate 2.0 1.5 Hydrogene peroxide 1.8 1.3 Permanganate (Na/K) 1.7 1.2 Chlorinie 1.4 1.0 Oxygen 1.2 0.9 Superoxide ion (O•) -2.4 -1.8 Source: ITRC 2005 [72].

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Hydroxyl radicals, OH •, react rapidly and unselectively with unsaturated or-ganic compounds, e.g. aromatic structures. The way in which OH • radicals are generated differs between different methods, but one of the most common ways is via iron-catalyzed decomposition of hydrogen peroxide at low pH (2.5-4.5). This reaction was discovered in the 1890s by H.J.H. Fenton and is known as Fenton’s reaction.

H2O2 + Fe 2+ → Fe 3+ + OH • + OH - (1)

At pH less than 5, the iron (III) is reconverted to iron (II) and the iron remains in solution, which is the reason for performing the reaction at low pH. Origi-nally, the Fenton reaction was also performed with a peroxide concentration of about 0.03 %. However, modified Fenton reactions are now commonly used during remediation with higher hydrogen peroxide concentrations, generally ranging from 4 % to 20 % for in situ treatments [72], at neutral pH, and with-out the addition of iron [73-78]. However, Tang and Huang (1996) [79] con-cluded that in addition to an optimal peroxide concentration there are also optimal peroxide and Fe2+ ratios (if iron is used). In the absence of organic compounds, excess ferrous ions are the primary scavengers of hydroxyl radi-cals:

Fe 2+ + OH • → Fe 3+ + OH - (2)

This means that if the ferrous ion concentration is too high, the reagent itself may consume the oxidant (reaction 2), with consequent reductions in oxidation efficiency. Hence, optimal conditions for Fenton oxidation should be carefully determined.

There has been great interest in remediation using Fenton reactions (reflected in the publication of large numbers of articles on many applications of this approach) due to their high oxidation power. Fenton reagent and Fenton-like applications have been extensively investigated and found to work well for the treatment of wastewaters [80, 81] and the remediation of both polluted soils and groundwater [73-77, 82-84].

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Ozonation is a common municipal water treatment technique. However, over the past 20 years increasing numbers of studies have also suggested that ozone can be successfully used to degrade complex organic contaminants in water and soil [72]. Unlike many other chemical remediation methods, the ozonation process involves the introduction of a gas, following which oxidation may pro-ceed either directly or indirectly. Direct oxidation occurs through reactions between the ozone molecules and the contaminants, while indirect oxidation involves the production of hydroxyl radicals. The relative importance of indi-rect oxidation increases with increases in pH, which enhance the production of hydroxyl radicals. Minerals present in the soil may also enhance the production of hydroxyl radicals, by providing catalytic surfaces for the hydroxyl radical-generating decomposition of the ozone molecules [85]. As mentioned earlier, hydroxyl radicals are non-selective and attack organic contaminants, preferen-tially unsaturated compounds, by breaking carbon-to-carbon bonds. Indirect oxidation with hydroxyl radicals is faster than direct oxidation, as indicated in Table 3, since hydroxyl radicals have higher oxidation strength than ozone. At high concentrations, ozone can act as a sterilizing agent. However, at lower concentrations, ozone has the potential advantage of introducing oxygen to the soil, which favours microbial activity for instance, and hence may be beneficial if bioremediation is to be used as a second remediation step [72].

Permanganate is a strong oxidation agent with high affinity for organic com-pounds containing carbon-carbon double bonds, aldehyde groups or hydroxyl groups. It can be applied in either of two common forms: sodium or potas-sium permanganate (NaMnO4 or KMnO4, respectively). Both have similar re-activity, but they are supplied in different states. The KMnO4 oxidation agent is prepared from a crystalline product with a maximum concentration of 4 %, while NaMnO4 is supplied as a solution with a concentration of 40 %, enabling the use of higher concentrations [72]. The feasibility of using permanganate oxidation has been investigated for treating contaminated groundwater [86], and both soil and sediments, often in combination with other treatments [87].

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Photolytic oxidation using ultraviolet (UV) light is a very common remediation method for groundwater treatment. To work well the turbidity and amounts of suspended solids in the water need to be low, allowing good transmission of light. Thus, photodegradation of dioxins, for instance, has been found to be negligible in soil, but they rapidly photodecompose when dissolved in ethanol [88]. Hence, UV treatment is unsuitable for treating soils directly, but photode-gradation is often used in conjunction with some solubility-enhancement method like solvent-washing, or extraction using vegetable oils or surfactants [57, 64, 89, 90]. It can also be used in combination with other oxidative meth-ods, using ozone or Fenton’s reagent for instance [80, 81, 91].

Pressurised hot water oxidation (PHWO) or supercritical water oxidation (SCWO), is also based on hydroxyl radical formation, but initiated by the introduction of oxidants under supercritical conditions. Air is the most commonly used oxi-dant, but other reagents include O2 and persulphate. Hydrophobic compounds in soil, with high chemical stability, can be oxidised by using SCWO [66, 67, 92, 93], since it is performed under strongly oxidative conditions. SCWO is said to be economically attractive for treating wastewater containing organic compounds at concentrations in the range 1-20 % [94], since the heat released during the oxidation will maintain the required reaction temperature, if prop-erly designed heat exchangers are used. However, at lower concentrations this technique is not economically attractive, and at higher concentrations of or-ganic compounds incineration is competitive. A drawback with these tech-niques is that they are relatively expensive with regards to the energy inputs and high demands on the equipment.

Reductive dehalogenation techniques Chlorinated contaminants may be detoxified through dechlorination under reducing conditions, possibly via a proposed mechanism involving nucleophilic substitution and oxidative dehalogenation of the halogenated compounds [95]. One such technique is the APEG (alkali metal hydroxide/polyethylene glycol) process, in which polyethylene glycol (PEG) is utilized as a reducing agent, often used to treat contaminants in soils, sludges, sediments and oils. The most commonly used hydroxide is potassium hydroxide (KOH) [47]. However, APEG may not be suitable for the remediation of dioxin-contaminated soils [96]. An alternative that may be faster and cheaper than APEG, and suitable for treating aqueous dioxin-contaminated soils, is base-catalysed decomposition (BCD). The BCD reaction requires hydrogen ions so a hydrogen donor is added, if no appropriate donors are present in the contaminated material. The process is performed at temperatures around 315-420 ˚C, with alkali addition from 1-20 % [97].

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3.3 Biological treatments

In biological treatments microorganisms (MO) are utilized to degrade hazard-ous organic contaminants, ideally to carbon dioxide and water. The contami-nants may be used as a food source by the MO, and biological treatments usu-ally involve stimulation of the MO in a site- and/or contaminant-specific man-ner. Biological treatments can take place under either aerobic or anaerobic conditions, and if pollutants are available to the organisms, it may be possible to use an in situ approach [98]. Traditional in situ biodegradation is usually used in combination with groundwater pumping and soil flushing to circulate nutri-ents and oxygen [47]. Organisms can also be introduced to the soil. However, Wilson and Jones consider in situ techniques to have limited effects on most PAHs [99], and as their availability declines other techniques may be required [98].

Landfarming is an above-ground remediation approach, in which contaminated soil is spread in a thin layer on the surface of the treatment area. In some cases MO are added and their activity is stimulated by the addition of nutrients, min-erals and moisture. The soil must be carefully mixed, to increase the contact between the organics and the MO, and to ensure there is sufficient oxygen to meet the demands of aerobic biological degradation. Landfarming has been shown to be effective for reducing concentrations of PAHs with three rings or fewer [99].

In composting naturally occurring MO are added, in a controlled environment, to compost made of contaminated soil to which nutrients, water, oxygen and material with high organic contents such as wood chips, straw or sawdust are added [47]. This results in higher concentrations of organic matter in the com-posting material, compared to those in landfarming material for instance. The high organic content promotes higher microbial activity and, hence, increases in temperature, which in turn result in higher degradation rates and more effi-cient degradation.

In bioreactors or bioslurries, the soil to be treated is mixed with water and aerobic biodegradation is performed under controlled conditions. Highly contaminated material can be treated in such biodegradation systems [47], which are regarded as being the easiest systems to manage and hence to maximise efficiency [98, 99]. Nevertheless, even these systems do not always successfully treat the more hydrophobic PAHs [99].

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To overcome limitations of biological treatments due to low bioavailability of the contaminants, combining them with chemical oxidation or separation techniques, using surfactants for instance, seems to be a promising option. Several studies report that this approach can increase the speed and efficiency of PAH removal [74, 75, 100-103]. Further, although only there have only been a few studies of the potential utility of bioremediation of dioxin-contaminated soils, combining biodegradation with pre-oxidation using Fenton’s reagent has yielded some encouraging results [104].

3.4 Thermal treatments

Incineration When incineration is used in the remediation of contaminated soils tempera-tures exceeding 1000 °C are usually applied. The high-energy demands make this technique very costly, although when soils with high organic contents are treated the energy gained from their combustion will help to reduce the total energy demands. Therefore, mainly highly polluted areas with total organic contents exceeding 20-25 % are likely to be considered for this kind of reme-diation [94]. Incineration is the most effective technique for degrading PAHs and chemically stable compounds such as PCDD/Fs.

Thermal desorption Thermal desorption can be considered a thermal separation technique, in which contaminated soils are heated to temperatures between 100 and 600 °C to vaporize contaminants with boiling points in that range. The vaporized con-taminants are not destroyed in this process, but they can be collected and de-stroyed by other methods.

Thermal desorption techniques can be divided into high and low temperature thermal desorption variants [105]. Low temperature thermal desorption works well on volatile compounds, while compounds with higher boiling point re-quire higher temperatures. Although many compounds will be oxidised rather than vaporised in the high temperature range, the process should not be con-sidered as an incineration process since destruction of the contaminants is not the intended result. However, as for incineration, thermal desorption is energy demanding and costly [47].

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Vitrification Stabilisation/solidification and vitrification are two groups of stabilisation tech-niques, both of which are particularly suitable for soils containing inorganic contaminants. The term stabilisation refers to the fact that contaminants in the treated soil are converted to less soluble, more immobile forms. However, the reason for mentioning this technique in this context is that the vitrification process involves use of extremely high temperatures, (1600-2000°C) and can thus be considered a thermal technique. During the high temperature processes of vitrification, organic compounds may be volatilised or pyrolysed [47, 106]. Hence, vitrification processes can be applied to solid matrixes contaminated with both inorganic and organic pollutants, for instance ashes, which often contain high levels of heavy metals, dioxins and PAHs. Vitrification of such ashes leads to reductions in the levels of the organic pollutants [107, 108], and stabilisation of inorganic contaminants. The soil is melted by the high tempera-tures and as it cools the material forms a hard, monolithic, chemically inert, stable glass and crystalline product with very low leaching characteristics [47]. However, it is important to remember that stabilisation techniques do not pro-vide a final solution to contamination problems, and another treatment tech-nique may be required in the future.

3.5 Innovative techniques

Nano-scale degradation In a recent review Zhang concluded that nanoscale iron particles offer potent possible options for environmental remediation [109]. Due to low operational costs, nanoscale particles have the potential to provide cost-effective remedia-tion solutions, although a major challenge will be to develop cost-efficient methods for synthesising the nano material [110]. The large surface areas and high surface reactivity of nanoscale particles make them highly suitable for remediation applications, with the ability to degrade a wide variety of common environmental contaminants, including chemically stable compounds such as dioxins [111]. The use of zero valent iron systems for nano-scale remediation is being extensively investigated and lately, the introduction of a second metal, i.e. bimetallic systems (zero valent iron and a catalytic noble metal, such as Pd or Ni [112]), has been found to significantly enhance dechlorination rates [111]. The mechanism responsible for dehalogenation when using zero valent iron alone is electron transfer, while in a bimetallic system the dehalogenation mechanism shifts to hydrogen atom transfer [111, 113].

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It should be noted that interest in nanotechnology is rapidly growing across several scientific disciplines, including electronics, biotechnology and the physical sciences [114]. Hence, the many areas for application leads to large amounts of nanomaterials that are being released into the environment are increasing exponentially. The fate and extent of exposure of nanomaterials to humans is not known as yet [110], however ironically some of the properties that make nano-scale particles attractive are also causes of concern from a health perspective [115]. Therefore, the environmental effects and extent of nanoscale exposure should be investigated in parallel with the development of nano-scale material.

Mechanochemical degradation Another recently developed technique for degrading organic pollutants is mechanochemical treatment. This technique involves grinding the polluted matrix in a ball mill at room temperature, and it has been shown to work for a wide variety or organic compounds, including dioxins [116] and PAHs [117]. To treat chlorinated compounds a dechlorinating agent such as CaO needs to be added. During grinding, mechanical energy is transferred from the milling bodies to the solid system, which is converted to thermal energy that enhances the surface reactivity and produces highly active radicals that may directly react with the pollutants [118]. The simplicity of the technology and the fact that no external heating is needed make this approach economically attractive and very promising for further development.

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4. AVAILABILITY OF POLLUTANTS IN SOIL

4.1 Bioavailability and availability

POPs present in the environment may pose profound risks to organisms since they may cause severe damage, due either to the inherently toxic properties of the parent compounds or through reactive intermediates, as discussed in Chap-ter 2. However, toxic POPs that are not available to the organisms cannot be considered to pose any risks to organisms. In other words, risks associated with POPs are highly dependent on their bioavailability, i.e. their accessibility or absorbability by living organisms.

Many organic contaminants become less available for organisms as they persist in soil, hence their uptake rates by organisms decline. This process often re-ferred to as ‘aging’. This phenomenon can be explained by sequestration of the contaminants in the soil matrix, involving processes such as covalent binding, sorption, diffusion and entrapment [119, 120] that are influenced by both the contaminants’ properties and the soil’s characteristics. The aging phenomenon has prompted discussion about whether risks posed by POPs are greatly over-estimated, since risk estimates are often based on total concentrations deter-mined by exhaustive extraction techniques [119, 120, 121] rather than bioavail-able concentrations. To overcome this problem, numerous attempts have been made to instead assess the bioavailability of organic pollutants in soil, by vari-ous means. For instance, in order to mimic the uptake of organic organisms, pollutants have been leached from soil or sediments using weak solvents [122-125], surfactants or semi-permeable membranes [126, 127]. However, in an assessment of the validity of these techniques Bergknut et al. [128] compared estimates of PAH bioavailabilities obtained using various techniques (solid phase microextraction, semi-permeable membrane devices, sequential leaching and leaching with different solvent mixtures and additives) to uptake rates by the earthworm Eisenia fetida as a reference system. The results showed that there were poor correlations between the results obtained by the tested tech-niques and the reference rates, therefore the authors concluded that mimicking biological uptake by chemical methods might be very difficult.

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Although the effect of sequestration is mostly considered from bioavailability perspectives, such as in risk assessments, mass transfer rates also affect overall degradation rates of contaminants sequestered in soils, since the contaminants become increasingly stable and resistant to remediation by both chemical and biological remediation methods. For instance, sequestrated contaminants show greater resistance to oxidative attack than contaminants in solution [76, 84, 129]. Therefore, the efficiencies of remediation techniques will vary depending on the treated soils’ characteristics, the contaminants’ properties and sequestra-tion time. This greatly complicates predictions of the success of specific reme-diation techniques, hence availability is as important as consideration for chemical remediation as it is for bioremediation and toxicological assessments. Therefore, in this thesis references to availability are not generally limited to either biological or chemical availability, but instead to broader, overall ‘avail-ability’.

4.2 Effects of soil characteristics on availability/ sequestra-tion

The sorption of contaminants to soil is a very complex process since soil is a highly inhomogeneous and complex matrix. Natural soil constituents, such as mineral particles and organic matter, are effective sorbents of contaminants sorbents, but residual products from former industrial activities may also con-tribute to the contaminant sorption [130]. Thus, their contributions to sorption processes are often considered in discussions regarding sequestration. In Fig-ure 5, a conceptual model of a soil particle is illustrated, together with the soil characteristics involved in the sorption/desorption processes. Soil organic mat-ter (SOM) is often regarded as the most important soil constituent for adsorp-tion of hydrophobic compounds [131-133]. When hydrophobic organic con-taminants are released into the environment they start to slowly migrate from the aqueous phase to the more hydrophobic surfaces or the organic soil frac-tion. This process leads to reductions in their availability for both chemical oxidation and biodegradation, and is highly dependent on the hydrophobicity of the contaminants [129, 134-140].

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SOM consists of particulate organic matter (POM) and dissolved organic mat-ter (DOM) and is very inhomogeneous. SOM may originate from natural or-ganic matter (NOM), such as dead plant parts and organisms, non-aqueous phase liquids (NAPL) and black carbon, all of which have varying adsorptive capacities. Black carbon is often present at sites such as former gasworks sites, and has been shown to adsorb hydrophobic compounds very efficiently [141, 142]. The physical structure of SOM and DOM has been shown to have a ma-jor impact on contaminant sorption [19]. Further, although DOM generally only constitutes a minor fraction of the total SOM, and hydrophobic com-pounds are mainly sorbed to POM [143], fine particulate matter and DOM generally make the largest contributions to the transport of organic pollutants from contaminated sites.

Figure 5. Conceptual model of a soil particle, indicating soil parameters involved in remediating activities and contaminant sequestration.

Contaminants in the soil matrix will first enter the macropores and sorb to the surfaces of the matrix aggregates [144], but with aging they will gradually mi-grate into the micropores, which are narrow, inaccessible apertures in and amongst the soil particles (Figure 5). Therefore, clay generally has negative effects on the degradability of contaminants in soil, since reductions in particle sizes increase the specific surface area, and thus increase the entrapment of

Fine structuresoil

Coarse soil

1. Bulk liquid/ oxidant/ solvent

Soot particles

Metal oxide coating

NAPL

Organicmatter

Mineral phase

2. Water film

3. Mesopore

4.Micropore

1.

2.

3.

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contaminants in narrow spaces. Hence, the contaminants will become much less available in clay soils. In addition, the clay structure may also affect the degradation efficiency. For instance, in some studies, PAH sorption has been found to be more strongly affected by clay minerals than SOM contents [145, 146], and Hwang and Cutright observed correlations between fractions of ex-pandable clay and both desorption and biodegradation [147]. While 82 % of pyrene was found to be degraded in soil with the lowest amount of expandable clay the cited authors tested, only 65 % was degraded in the soil with the high-est amount of expandable clay. This can possibly be explained by exposure of large available internal surface areas for sorption during expansion of the clay [148].

The sequestration of one compound will also probably be affected by the pres-ence of other compounds, and the order in which they were introduced [136], since sorption sites may be saturated by compounds or groups of compounds added in early contamination stages. In such cases the availability of com-pounds deposited later will probably be greater than that of previously depos-ited compounds. However, many researchers still study the bioavailability of aged PAHs using relatively low spiked concentrations [135], even though bio-degradation parameters are known to be to be influenced by numbers of co-contaminants [149]. Higher percentages of pollutants are generally degraded in highly polluted soils [136, Paper I] and high concentrations of contaminants are often present in field samples.

Another important variable of soils is their oxide content. Oxides have been shown to promote the formation of hydroxyl radicals during ozonation [85], which should increase the treatment efficiency since hydroxyl radicals are more reactive than ozone molecules. However, the effects of the types and amounts of oxide present in a soil have been mostly considered in the literature in the context of Fenton reactions, since iron is normally added as a catalyst in Fen-ton reaction treatments. Naturally occurring oxides may also serve as catalysts in Fenton reactions, according to results presented in numerous studies [ 76, 77, 150], and data presented by Watts et al. suggest that Fenton-like reactions are catalysed on the surfaces of iron minerals in the presence of H2O2 [76]. The use of iron oxide has also been found to be effective at neutral pH. Hence, use of iron oxides rather than standard Fenton’s reagent has several advantages; notably they can be more easily used for in situ applications since there is no need for additions of FeSO4, nor pH adjustments [151].

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Despite the common use of iron in the classic Fenton reaction, which involves the oxidation of Fe (II) to Fe (III), and iron minerals as natural catalysts in Fenton-like reactions, Miller and Valentine have postulated that manganese oxides may be even more important catalysts in some cases [152, 153]. In ad-dition, under highly oxidising conditions amorphous iron oxyhydroxides may participate in oxidative reactions on surfaces in systems containing H2O2 at high concentrations, and in such cases high H2O2 additions may be required for effective decontamination [77, 154].

In summary, soil characteristics affect not only the sequestration of com-pounds in the soil, they may also affect the efficiency of the selected remedia-tion method to various degrees. This clearly demonstrates the complexity of sorption/desorption processes.

4.3 Aged soils and spiked soils

A distinguishing characteristic of soils at former industrial sites is their hetero-geneity. In addition to variations in the composition of the original soil, the contamination resulting from the former industrial activities varies greatly. For that reason, controlled experiments are often considered to be too complicated to perform on aged industrial soils [155]. Instead, many researchers use spiked soils in order to simplify degradation or availability studies. A major advantage of using spiked soil is the possibility it provides to control the parameters of interest. However, significant differences have been observed in the behaviour of contaminants recently added to soil samples, even if they are aged in the laboratory, compared to that of contaminants in weathered and field-contaminated soils [78, 119, 155, 156]. Further, just one or a few contaminants have been used in many studies, while real environmental samples, as men-tioned above, often contain complex mixtures of both pollutants and other compounds. Therefore, spiked soils should be used with great caution, if at all, in remediation or availability studies, and ideally multiple parameters should be considered, otherwise the results may not be representative of true remedial conditions, and there is also a great risk that important interaction effects may be overlooked.

30

To overcome some of the problems that are often associated with experiments performed with aged soils, multivariate analysis may be a useful tool, e.g. Par-tial Least Squares (PLS) multivariate regression, which enables the simultane-ous evaluation of the effects of many parameters. Two matrices are considered simultaneously: a descriptor matrix and a response matrix, and components that describe the variations and relationships between the matrices are calcu-lated. PLS modelling is capable of handling noisy and missing data [157] and evaluating large numbers of parameters, simultaneously. For instance, in Paper II, ten aged soils were characterised with regards to >15 soil parameters and the concentrations of 24 PAHs were determined. The influence of the soil parameters and physicochemical properties of the PAHs on their chemical degradation by Fenton’s reagent in the soils was then investigated by PLS modelling, which detected significant interactions between degradation effi-ciency and the soil characteristics and PAHs’ physicochemical properties.

31

5. APPLICABILITY OF SOIL REMEDIATION METHODS

As discussed in Chapter 3, there is a wide variety of remediation technologies to choose from, and the effectiveness of many of them are highly site- and contaminant-specific. Hence, site characterisation is a critical step, especially if in situ remediation is to be applied, since thorough knowledge of the contami-nant profile and site characteristics is crucial in order to select the most appro-priate technique. For that reason it is of utmost importance to learn not only more about the parameters influencing the sequestration and treatability of aged, contaminated soils, but also how individual remediation methods are affected by these parameters. In the studies underlying this thesis, several remediation methods (chemical and physical) were applied to a wide variety of aged, contaminated soils. The degradation and removal of PAHs and dioxins from these soils were evaluated and the results obtained were used to assess the relationships between the soils’ characteristics (including their contamina-tion histories), the contaminants’ physico-chemical properties and the effec-tiveness of the applied treatments.

For clarification, in the discussion throughout this thesis, low molecular weight (LMW) PAHs refers to PAHs with two and three fused aromatic rings, whereas high molecular weight (HMW) PAHs refers to PAHs with five and six aromatic rings. PAHs with four rings have chemical properties that are similar to both groups and are therefore not included in any of the groups.

32

5.1 Remediation methods applied to aged soils

Chemical oxidation methods [Papers I & II] Fenton and ozone are two of the most commonly investigated oxidation tech-niques [158, 159]. Both methods are based on the formation of hydroxyl radi-cals and have been shown to degrade PAHs effectively. In Papers I and II, these two chemical oxidation methods were applied to several aged, PAH con-taminated industrial soils under varying conditions

In Paper I, Fenton’s reagent was added with a hydrogen peroxide concentra-tion of 24 % and elevated temperature (70 ˚C), which induced a vigorous reac-tion. In addition, ozone oxidation was carried out in a solution of 50 % etha-nol. In Paper II the soils were subjected to Fenton’s reagent under milder oxidative conditions than in Paper I; the hydrogen peroxide concentration was 5 % and the reaction was performed at 25 ˚C. In this thesis the Fenton reagent concentrations used in Paper I are referred to as “strong” whereas the concen-trations used in Paper II are referred to as “weak”. The PAH degradation effi-ciencies of the tested chemical oxidation methods were investigated and com-pared. The soil samples included in Papers I and II, described in Table 4, originated from sites associated with three different kinds of industrial activi-ties (gas works sites, wood-impregnation sites and a coke production site that was still in use when the samples were collected). The activities at the sites started as early as 1893 at one site, thus the contaminants had had substantially varying times to become sequestered into the soils (Table 4).

The degradation efficiencies of the Fenton and ozone treatments used in Pa-per I in the tests with nine different soils ranged from 40 to 86 % and from 10 to 70 %, respectively (Figure 6A). The degradation efficiency was significantly lower (9-43 %) when the weaker Fenton’s reagent was applied in Paper II. There may have been several contributory reasons for the between-soil varia-tions, including differences in the degree of contaminants’ sequestration in the soils and soil characteristics, since the investigated soils varied from fine sand to coarse sand, and also included a sediment sample. In addition, differences in the soils’ origins probably affected the degradation efficiency, since they will affect the soils’ contamination patterns. In most soils PAHs with two, three and four fused aromatic rings account for more than 90 % of the total con-tents with a few exceptions (Figure 6B); The concentration of two through four aromatic rings only accounted for 70 % in Site 7. The soil from that site consisted of coarse sand (Table 4), which may have facilitated leaching of smaller and more water soluble PAHs from that site (LMW PAHs in particu-lar). In samples from the gasworks and coke production sites ( Soil 4, 9 and 10)

33

there are higher concentrations of PAHs with five and six aromatic rings (16-26 %) than in the major part of the wood impregnation sites (generally 10 %) (Figure 6B). Several reasons may have contributed to these different contami-nation profiles. The different contamination patterns may have been source-related and/or, the abundance of smaller PAHs may have declined with time at the gasworks site, since contamination of that site started in the 19th century. Hence environmental processes, i.e. natural degradation and evaporation, may have caused a depletion of smaller PAHs from soil sample 9 and 10. Further-more, a relatively even ring distribution of concentrations of PAHs with two through six aromatic rings, is shown in soil sample from Site 4, the coke pro-duction site, which was still in use when the soil sample was collected. Hence, this even ring distribution pattern may be an indication of a “young” PAH profile (Paper I).

Figure 6. (A) Degradation efficiencies of PAHs achieved in tests with aged soils using strong Fenton’s reagent (Fenton high, Paper I), ozone (Ozone, Paper I), and weak Fenton’s reagent (Fenton low, Paper II). W, C and G indicate soils from a wood-impregnation site, coke production site, and gasworks site, respectively. (B) Relative abundance of two through six aromatic rings prior to treatment with weak Fenton’s reagent in Paper II.

0102030405060708090

100

1(W)

2(W)

3(W)

4(C)

5(W)

6(W)

7(W)

8(W)

9(G)

10(G)

%

Ringdistribution

2 Rings 3 Rings 4 Rings 5 Rings 6 Rings

0102030405060708090

1(W)

2(W)

3(W)

4(C)

5(W)

6(W)

7(W)

8(W)

9(G)

10(G)

%

Degradation ef f iciency

Fenton (I) Ozon (I) Fenton (II)

(A) (B)

34

Table 4. S

ite

loca

tions

, ind

ustr

y ch

arac

teri

stic

s an

d so

il ty

pes

of s

ampl

es in

clud

ed in

Paper I and II.

Site Location

Coordinates

Industry

Period of

operation

Sampling-

depth

Soil

Observations

Paper

1

Holmsund

N 63º 42´ E 20º 21´

Wood preservation

1943-1983 20-30 cm Sandy Till

Black aggregated soil, strong smell of tar

I & II

2

Holmsund

N 63º 42´ E 20º 21´

Wood preservation

1943-1983 10-20 cm Sandy Till

Black aggregated soil, strong smell of tar

I & II

3

Holmsund

N 63º 42´ E 20º 21´

Wood preservation

1943-1983 10-20 cm Sandy Till

Smell of tar

I & II

4

Luleå

N 65º 35´ E 22º 09´

Coke production

a

Top soil

Sediment

Black waterlogged sediment, oily film

on

water surface

I & II

5

Forsmo

N 63º 16´ E 17º 12´

Wood preservation

1933-1950 2-18 cm

Fine sand

Black aggregated well sorted sandy soil

I & II

6

Forsmo

N 63º 16´ E 17º 12´

Wood preservation

1933-1950 0-10 cm

Fine sand

Black aggregated well sorted sandy soil

I & II

7

Hässleholm N 56º 09´ E 13º 46´

Wood preservation

1946-1965 40 cm

Coarse sand Smell of tar

I & II

8

Hässleholm N 56º 09´ E 13º 46´

Wood preservation

1946-1965 40-60 cm Coarse sand Smell of tar

I & II

9

Husarviken

N 59º 21´ E 18º 06´

Gas work

1893-1972

b

Sand

Smell of tar

I & II

10

Husarviken

N 59º 21´ E 18º 06´

Gas work

1893-1972

b

Sand

Smell of tar

II

35

Solubility enhancement by low molecular weight alcohols To enhance the availability of contaminants in highly weathered soils, oxida-tion treatments sometimes include the use of surfactants or water-soluble sol-vents, such as low molecular weight alcohols [160, 161], particularly to desorb lipophilic compounds. For instance, Bogan et al. found that pre-treatment with vegetable oils had positive effects on the chemical degradation of PAHs [65], mainly for HMW PAHs, which are normally strongly sorbed to soil organic matter [65]. However, due to the non-selective nature of hydroxyl radicals, they may react with the dissolving media [160], thereby increasing the amounts of oxidant required, and reducing the mineralization rates, compared to using chemical oxidation alone [65, 101, 103, 162].

Degradation efficiency was similar for PAHs with 2 through 6 rings (53-72 %), with Fenton treatment in Paper I, where no ethanol was used. The overall degradation efficiency was nearly two magnitudes higher with the strong Fen-ton’s reagent (62 %) than with ozone treatment (32 %). With the ozone treat-ment in Paper I, significantly lower degradation of HMW PAHs and PAHs with four rings were noted, than LMW PAHs (11-17 % and 33-44 % respec-tively) despite the use of ethanol. Hence, it is not possible to identify any clear effect on the degradation efficiency of HMW PAHS from the use of ethanol in Paper I (Figure 7A).

Nevertheless, screening experiments were conducted on Soil 10, prior to the ozone treatment in Paper I and the results from the screening experiments showed that ethanol has a substantial positive effect at a concentration of 90 % on the overall degradation efficiency (Figure 7B). This indicates that contami-nant desorption may be a more important limiting factor under the selected conditions in Paper I than the suggested consumption of hydroxyl radicals during chemical oxidation. Furthermore, the increased degradation efficiency was particularly pronounced for PAHs with four to six aromatic rings, for which the degradation efficiency was increased from virtually zero to 44 - 67 %, as the ethanol concentration was increased from 10 to 90 %. Increases in time and temperature also had positive effects on the degradation efficiency. However, despite the increased degradation efficiency with higher ethanol con-centrations, it was limited to 50 % in assessments of treatment time and tem-perature, since ethanol concentrations above 50 % were associated with evapo-ration problems at elevated temperatures.

36

Selection of oxidant concentration

Addition of excess H2O2 and vigorous oxidation conditions have been docu-mented to increase the degradability of strongly sorbed contaminants [77, 165], possibly because they reduce the strength of sequestration. The similar degra-dation efficiencies of LMW PAHs and HMW PAHs that were achieved when the strong Fenton’s reagent was applied in Paper I (Figure 7A), supports this hypothesis.

When the Fenton’s reagent was applied under milder conditions, in Paper II, the overall degradation was only 14 % (compared to 62 %, with the stronger Fenton’s reagent in Paper I) and there were also substantial differences in deg-radation efficiency between LMW and HMW PAHs when Fenton’s reagent was applied in Paper II. Hence, increasing the oxidant concentration increases the degradation efficiency [165]. However, optimisation of the oxidant load is important since the opposite effect, i.e. reductions in degradability, may be observed if too high oxidant loads are used, due to self-consumption of the oxidant [79, 87] (as explained in section 3.2).

Figure 7.

(A) Degradation efficiency of two- to six-ring PAHs using Fenton’s reagent and ozone, observed in Paper I (nine soils)

(B) Ozone degradation during screening experiments with soil from Husarviken ( Soil 10), and high (90 %), medium (50 %) and low (10 %) ethanol (EtOH) con-centrations. Screening results from one experiment (with high ethanol concentration, low tem-perature and short time) were excluded due to analytical prob-lems.

0

10

20

30

40

50

60

70

80

90

1 2 3 4 5 6 7

(%)

Number of Rings

Low EtOH Medium EtOH High EtOH

0

10

20

30

40

50

60

70

80

1 2 3 4 5 6 7

(%)

Number of Rings

Fenton Ozone

(B)

(A)

37

Physical remediation methods [Papers III & IV]

Solvent washing – selection of the experimental domain

During solvent washing it is important to use a liquid that is capable of dissolv-ing the contaminants to be removed from the soil matrix. Therefore, the sol-vent must be carefully selected. In Papers III and IV ethanol was chosen, based on the results (unpublished data shown in Figure 8) from a prestudy performed prior to the experiments presented in Paper III. Extraction effi-ciency of dioxins and furans was investigated by comparing removal efficiency from a sandy soil by three solvents (methanol, ethanol, and a mixture of methanol and iso-propanol), using an accelerated solvent extraction (ASE) system. The overall removal efficiencies were compared with toluene in an ASE system at 150˚C and Soxhlet extraction. The results showed that all three of the investigated solvents were effective under the selected conditions and had similar removal efficiencies of PCDD/Fs (85-102 %), but ethanol was selected for further experiments due to its relatively low toxicity and low health impacts. Similar comparisons have been made previously with similar out-comes, i.e. ethanol has been chosen as the most suitable solvent for dissolving hydrophobic contaminants, either as a sole solvent or in combination with other co-solvents [54, 56].

Figure 8. Overall efficiency of extraction of PCDDs and PCDFs from samples of a sandy soil (from Hansson sawmill site), using various solvents and an accelerated sol-vent extraction (ASE) system.

0

20

40

60

80

100

120

40 ̊ C 55 ̊ C 40 ̊ C 65 ̊ C 40 ̊ C 65 ̊ C 150 ̊ C

Methanol Ethanol 2-propanol and

methanol

Toluene Soxhlet Zero

Ex

tra

cti

on

eff

icie

nc

y (%

)

38

After the solvent had been selected, screening experiments were performed to determine the experimental domain for further investigations regarding solvent washing of PCDD/F-contaminated soils. Screening experiments were per-formed using Soil 1 in Table 5, a sandy soil from a former samill site in the northern Sweden. Parameters investigated were the solvent concentration, temperature and washing time. The results indicated that the extraction effi-ciency was mainly dependent on the ethanol concentration, and the tempera-ture and washing time were of minor importance [Paper III]. The high de-pendence on the ethanol concentration is largely due to the very low water solubility of 2, 3, 7, 8-substituted dioxins and furans. The extraction of more readily water-soluble compounds seems to be less dependent on the ethanol concentration, for instance Khodadoust et al. [166] found that 50 % etha-nol/water and more concentrated ethanol solutions remove PCP from soils equally efficiently.

Although the maximum tested ethanol concentration (95 %) gave the highest removal efficiency, a solvent strength of 75 % ethanol was chosen for subse-quent investigations reported in Paper III, in order to avoid overestimating the extraction capacity when washing soils with high water contents. The soils that were investigated in study III are presented in Table 5. The soils differed with respect to PCDD/F concentrations as well as soil structure.

Application of solvent washing to aged soils

The efficiency with which the solvent-washing process removed PCDD/Fs in tests with Soil 1 ranged from 15-56 % during the screening experiments [Pa-per III]. However, the use of ten washing cycles greatly improved the results, increasing the removal efficiencies in tests with Soils 1 and 2-4 to 81 % and 85-98 %, respectively. Hence, after ten washing cycles the overall removal effi-ciencies did not vary significantly between the aged soils, despite the differ-ences in their origins. However, there were substantially greater between-soil differences in removal efficiency after the first washing cycle in tests with the same four soils (18-36 %), probably due to the differences in characteristics and origins of the four soils. Nevertheless, by applying several washing cycles acceptable remediation results were achieved for a variety of aged soils, in ac-cordance with observations in several previous studies [56, 58, 70, 167, 168, 169]. However, the highest amounts of PCDD/Fs were removed during the first three to five washing cycles, (Figure 9), indicating that fewer washing cy-cles could be used with almost as good results. Hence, by determining an ac-ceptable end-point for the remediation, the number of washing cycles, and thus costs, could possibly be reduced.

39

Figure 9. Amounts of PCDD/F (pg TEQ g -1 dry weight) in the extracts (x-axis), from solvent extraction cycles 1, 2, 3, 5 and 10 obtained during the sequential extrac-tion (y-axis) of Soils 1 - 4 at 60 ºC, with exponential trend lines fitted to the data (Pa-per III).

Optimization of granular activated carbon (GAC) adsorption

During the washing of aged soils, the dissolved contaminants are transferred to the washing fluid, and these highly toxic by-products need to be post-treated in some way. The most appropriate post-treatment depends on the contaminants present in the washing fluid. Activated carbons (AC) have excellent capacities to adsorb a variety of organic contaminants [170] and are extensively used for decontaminating polluted water and air. For instance, in a study by Ayotamuno et al. [171], removal efficiencies of 96 and 99.9 %, of petroleum from contami-nated groundwater were obtained with granular and powdered activated car-bon, respectively. Powdered AC has larger surface areas, and thus more ad-sorption sites, higher adsorption capacities and greater removal efficiency than granular activated carbon (GAC). Hence, use of powdered AC was recom-mended in the cited study [171]. However, it may cause clogging problems. Therefore, a commercially available GAC with high permeability, suitable for liquid applications, was selected for investigations of the potential utility of activated carbon as a PCDD/F-adsorbent for regenerating ethanol washing solvents in Paper IV. In this study batch-wise adsorption experiments were performed.

-

1

2

3

4

5

6

0 2 4 6 8 10

Soil 1

-

100

200

300

400

500

600

0 2 4 6 8 10

Soil 2

-

100

200

300

400

500

600

700

800

0 2 4 6 8 10

Soil 3

-

500

1 000

1 500

2 000

2 500

3 000

3 500

0 2 4 6 8 10

Soil 4

40

Table 5. L

ocat

ion

of s

ampl

ing

site

s, in

dust

rial

cha

ract

eris

tics

, soi

l typ

e an

d so

il pa

ram

eter

s of

sam

ples

use

d in

Papers III and IV

.

Soil

Location

Coordinates

Contaminating

activity

Soil description

LOI

%

Water

content

%

Concentration

pg-TEQ g -1 dry

weight

Paper

1

Hanssons sawmill,

Luleå

N 63º 35´

E 22º 09´

Wood preservative

site 1961-1975

Sandy-silty

3.4

5.9

30

III

2

Hanssons sawmill,

Luleå

N 63º 35´

E 22º 09´

Wood preservative

site 1961-1975

Sandy-silty

3.4

5.9

950

III & IV

3

Öbacka waterside,

Umeå

N 63º 50´

E 20º 15´

Wood preservative

site 1960`s-1975

High concentration

plant parts/organic

matter

40

51

2300

III & IV

4

Eka Chemicals,

Bohus

N 57º 51´

E 12º 01´

Industrial site

Fine particle clay-silty

soil, graphite particles

3.7

15

8100

III

41

The influence of organic content in the washing fluids on GAC adsorption parameters was investigated in tests with fluids used to wash samples of the sandy silty soil and the soil with high organic contents used in Paper III. The adsorption curves obtained show that the contaminants were efficiently re-moved from the solvents under the given conditions (Figure 10), although the characteristics of the two soils differed greatly. Hence, the investigated acti-vated carbon possesses the ability to adsorb PCDD/Fs. However further, pi-lot-scale investigations are needed before full-scale applications could be con-sidered.

Figure 10. Adsorption of PCDD/Fs (%) from fluids used to wash soils with low and high organic contents. .

-3

17

37

57

77

97

0,0 0,2 0,4 0,6 0,8 1,0 1,2 1,4 1,6

PCDD/F Adsorbed (%)

Carbon (g)

Low organic soil

-3

17

37

57

77

97

0 0,2 0,4 0,6 0,8 1 1,2 1,4 1,6

PCDD/F Adsorbed TEQ (%)

Carbon (g)

High organic soil

42

Table 6. C

hara

cter

isti

cs o

f so

il sa

mpl

es in

clud

ed in

Paper II.

Soil number

1

2

3

4

5

6

7

8

9

10

Characteristics

Age

(years)

40

40

40

- 62

62

48

48

71

71

Clay

(%)

0.00

0.30

0.20

0.40

0.00

0.00

0.00

0.00

0.20

a

Silt

(%)

5.6

4.9

7.4

15

4.8

36

3.7

0.1

7.6

a

Sand

(%)

94

95

92

85

95

64

96

> 99

92

a

Spec surface

(m2 /m

3 soil)

1100

1100

1200

1600

1000

2600

980

800

1200

a

LOI

(%)

9.8

6.9

2.7

13

2.6

13

6.8

2.2

19

12

pH

(H2O)

6.9

7.2

7.5

11

7.9

7.6

7.2

7.5

7.8

a

pH

(CaCl 2)

6.9

7.5

5.1

11

4.8

4.7

6.6

6.7

7.1

a

Cond

(mV)

5.0

-11

-30

-230

-98

30

-12

-28

-50

a

Al C

D

(g/kg)

0.40

a

a

0.50

0.20

1.8

0.40

0.10

0.60

a

Al O

x (g/kg)

0.50

a

a

1.1

0.17

2.0

0.46

0.24

1.4

a

Fe CD

(g/kg)

2.0

a

a

15

1.2

5.3

7.4

0.60

23

a

Fe Ox

(g/kg)

1.1

a

a

6.3

0.44

4.2

4.3

0.18

18

a

Mn CD

(g/kg)

0.10

a

a

0.20

0.00

0.10

0.70

0.00

0.20

a

Mn Ox

(g/kg)

0.05

a

a

0.00

0.00

0.08

0.34

0.00

0.09

a

a so

il ch

arac

teri

stic

s n

ot d

eter

min

ed

43

5.2 Influence of soil characteristics’ and contaminants’ physico-chemical properties on availability/sequestration

Chemical remediation methods [Papers I & II] In Paper II the effects of the soil characteristics and PAHs’ physico-chemical properties on their degradability in treatments with Fenton’s reagent were evaluated using multivariate data analysis. Information regarding the origins of the ten aged soils used in Papers I and II is included in Table 4. The ten soils were characterised with respect to the soil parameters shown in Table 6. The soils were of varying age and structure, as illustrated by the differences in their specific surfaces. However, the differences in specific surface between the soils were mainly due to variations in their sand and silt content, since they all gen-erally had very low amounts of clay material (less than 0.5 % in all soils, Table 6).

Aging and sequestration are known to be influenced by soil characteristics, origins, and contamination patterns. Hence, the characteristics of the ten soils included in the studies probably contributed to the variations in degradation efficiencies described in Papers I and II. However, relevant correlations are difficult to identify simply by visually examining raw data obtained for a num-ber of soils, instead an appropriate multivariate data analysis tool may be re-quired, as described below.

Influence of soil characteristics

In Paper II, PLS modelling was used to explore the possible relationships be-tween soil characteristics and degradation efficiencies. Results from the PLS analyses of the influence of different soil characteristics on PAH-degradability (using Fenton’s reagent) are shown in Figure 11. The response data were the degradation efficiencies in the ten soils. The PLS calculations yielded two sig-nificant Principal Components, explaining 81 % of the variation in the data and a predictability (Q2) of 24 %. The results (PC1 versus PC2 loading plot) illus-trated in Figure 11 are discussed in more detail below, but first the main fea-tures that can be seen in the figure will be briefly summarised. The PAHs are located in two separate groups in the loading plot, consisting of HMW PAHs and two to four aromatic rings, indicating that soil parameters affect the de-gradability of these groups to differing degrees.

44

-0,40

-0,30

-0,20

-0,10

0,00

0,10

0,20

0,30

0,40

-0,40 -0,30 -0,20 -0,10 0,00 0,10 0,20 0,30 0,40

PC2

PC1

%LOI

2 Ring

3 Ring

4 Ring

5 Ring

6 Ring

Age

Al CD

Al Ox

Fe CDFe Ox

LOIMn CD

Mn Oxcond

pH_H20

spec_surface

The figure also shows that loss on ignition (LOI), oxide content and specific surface are located on the opposite side from PAHs with less than five rings, indicating that these parameters have a negative effect on the degradation effi-ciency of these PAHs, i.e. increases in the LOI, for instance, are associated with reductions in their degradation efficiency. For HMW PAHs, age and pH are negatively correlated, whilst conductivity and the amount of degraded or-ganic matter (%LOI) are positively correlated to the degradation efficiency (Paper II).

Increasing the organic matter content apparently favoured the sorption of LMW PAHs and PAHs with four rings to the soils, since their degradation was reduced with increases in LOI values. In contrast, the degradation of HMW PAHs was not significantly affected by increases in organic matter contents. Hence, it can be concluded that sufficient organic matter was present in the soils with the lowest LOI values (2.2 %) to adsorb the HMW PAHs very strongly, and that increases in LOI did not increase the sorption strength any further for these compounds. (Figure 11)

Recent studies have suggested that the degradation of 3-6 ring PAHs is af-fected by the total organic carbon (TOC) content in soils with TOC >5 %, while for soils containing less than 5 % organic matter the LMW PAHs have a

Figure 11. Results (PC1 versus PC2 loading plot) of PLS calculations of the influ-ence of measured soil characteristics on PAH degradation efficiency with Fenton’s reagent (Paper II).

45

stronger dependency on soil porosity [129]. The results in Paper II conflict with these suggestions, and show that organic matter had a stronger effect than the specific surface on the sequestration of PAHs containing less than five fused rings.

For reasons that will be explained in detail below, the most likely reason for the declining degradability of LMW PAHs with increasing amounts of SOM seems to be that contaminants were strongly sorbed to the SOM, which may be viewed as sorbents in which the hydrophobic contaminants migrate and be-come less accessible with time, i.e. in which sequestration of the contaminants will occur [129]. Another common postulation is that the non-selective hy-droxyl radicals react with the soil organic matter instead of the contaminants, thereby inhibiting the process due to consumption of the oxidant. In accor-dance with this hypothesis, Lindsey and Tarr have found that reactions of hy-droxyl radicals with non-pollutant compounds can reduce degradation efficien-cies [172, 173]. However, this explanation is not likely to be applicable to the results reported in Paper II, since the amount of degraded organic matter was found to be positively correlated to the degradation of PAHs with five and six rings (i.e. there was a clear positive correlation between the degradation of HMW PAHs and % LOI).

46

In Figure 12, a proposed desorption/oxidation mechanism for HMW PAHs sorbed to SOM is shown. The oxidation process most likely caused the release of smaller parts of organic matter from larger molecules, reducing the hydro-phobicity of the natural organic matter, and the oxidation of SOM resulted in the release of sorbed contaminants [174], which were then further oxidised by the excess of hydrogen peroxide.

Figure 12. Proposed desorption/oxidation mechanism of HMW PAHs sorbed to SOM with increasing availability of organic contaminants as a result of soil organic matter degradation [Paper II].

Figure 13 shows results from the PLS calculations of the influence of the PAHs’ physico-chemical properties on the observed PAH degradability in Pa-per II. Mean degradation efficiencies of each of the 24 PAHs in the ten soil samples were used as response data. The PLS calculations yielded two signifi-cant Principal Components, explaining 81 % of the variation in the data with a predictability (Q2) of 56 %. Physico-chemical parameters reflecting the size and hydrophobicity of the PAHs were located to the left in the loading plot and were negatively correlated to the degradation efficiency (Fenton) in the soils. Correspondingly, PAHs consisting of five and six rings had low PC1 scores.

Soot particles

Metal oxide coating

NAPL

Organic matter Mineral phase

• OH

Sorbed unavailable organic contaminant

Increased availability of organic contaminant as a result of oxidized SOM

Oxidation of soil organic matter

47

Smaller PAHs, with two or three rings are more water soluble (Table 1) and had higher PC1 scores. Due to their higher water-solubility they become in-creasingly available to the oxidant, as reflected in the positive correlation be-tween degradation efficiency and log Sw. The second PC indicates the second most important parameters for degradability. The ionisation potential (Ip-homo), which reflects the chemical reactivity of the compounds, and five-ringed PAHs (5 ring) were located in the upper part of the loading plot, and in this model these parameters were also positively correlated to degradation effi-ciency (Figure 13).

Soil-sorbed HMW PAHs are generally more resistant to degradation than LMW PAHs. However, surprisingly, perylene and benzo(a)pyrene were re-moved almost as efficiently as phenanthrene and anthracene in Paper II, probably because of the relatively high ionisation potentials (which is influ-enced by the ring configuration) of these HMW PAHs. The importance of the ring configuration on degradability is illustrated by the data for phenanthrene and anthracene. In Paper II anthracene was more efficiently degraded than phenanthrene, in accordance with findings reported by Lundstedt et al. [160], but Nam et al. found the opposite, i.e. that anthracene had higher oxidation

-2

-1

0

1

2

-12 -10 -8 -6 -4 -2 0 2 4 6 8 10 12

t[2]

t[1]

2

3

3

3

3

3

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4

445

555

5

5

66

-0,40

-0,20

0,00

0,20

0,40

0,60

0,80

-0,20 -0,10 0,00 0,10 0,20

w*c[2]

w*c[1]

5 ring

Ea-lumo

Fenton

Hvap

Ip-homo

L/B

MPNOR

TSA

log Koclog Sw

m.vol.

Area log Kow

BP Mw

F NOAR

HOMO Pol

LUMO Ref

Vol

Figure 13. PLS score (top) and load-ing (bottom) plots (PC1 versus PC2), showing the influence of PAHs’ phys-icochemical properties on their chemical degradabil-ity with Fenton’s reagent. Symbols are explained more into detail in Paper II.

48

resistance than phenanthrene [78]. Lundstedt et al. suggested that this could have been due to ethanol pre-treatment, since anthracene dissolved in ethanol has been shown to be highly susceptible to oxidation by Fenton’s reagent [160]. However, since no such pre-treatment was applied in Paper II there may be an additional explanation for this discrepancy. Phenanthrene has higher water solubility, which is generally associated with higher availability and de-gradability. On the other hand, phenanthrene has an angular configuration, which generally provides more chemical stability than the linear configuration of the anthracene molecule [8] (Table 7).

The results presented in Paper II indicate that the chemical reactivity of phe-nanthrene had a more pronounced effect on its degradability than its water solubility. This conclusion is further supported by the degradation results for perylene and benzo(a)pyrene, which are also known to have high reactivity towards hydroxyl radicals. These compounds were more or equally efficiently removed as phenanthrene and anthracene by all three oxidation methods used in Papers I and II. Similar results have also been reported by Lundstedt et al. and Nam et al. [160, 78].

Table 7. Ring configuration and a selection of physico-chemical properties of six PAHs

PAH Ip-homo Log Sw Mw

Acenapthene

-8.495 0.029 154

Fluorene

-8.711 0.012 166

Anthracene

-8.123 0.00037 178

Phenanthrene

-8.617 0.0072 178

Perylene

-7.858 1.2E-05 252

Benzo(a)pyrene

-7.922 1.5E-05 252

49

Physical remediation methods [Papers III & IV] The sampling locations and characteristics of the contaminated soils used in Papers III and IV are presented in Table 5. The soils had very different char-acteristics, e.g. their organic contents, in terms of LOI values, varied between 3.4 and 40 %, while their PCDD/F concentrations ranged from 30 to 8100 pg-TEQ g-1 dry weight. Soils 1 and 2 had a sandy character with low organic con-tents (3.4 %), while Soil 3 originated from a former sawmill site and had large quantities of wood fibres, typical of soils from such sites (LOI 40 %). Soil 4, which was a clay soil, also contained graphite particles.

Influence of soil characteristics [Paper III]

In Table 8 total removal efficiencies are shown for Soils 2-4 after ten washing cycles at 60 ˚C. As discussed above, the total extractability of the contaminants in the four soils after ten washing cycles was quite similar, ranging from 81 to 98 % (Table 8). To acquire more information regarding the extraction process during solvent-washing, the fraction of PCDD/Fs removed from the soil through particle separation was estimated. This was done by first calculating the dissolved fraction from the extraction curves for each soil, which were based on the PCDD/Fs found in the filtrates (Figure 9). This amount was then subtracted from the total amount of PCDD/F removed. The difference be-tween these two amounts provides an estimate of the fraction of PCDD/Fs sorbed to small particles that were subsequently filtered off, and hence re-moved.

The estimates of particle removal efficiency varied more than the total removal efficiency. For instance, particle removal efficiencies for Soils 1 and 2, which had similar characteristics, were 50 and 16 %, respectively (Table 8). Since Soil 2 contained much higher PCDD levels than Soil 1 it is likely that the fine frac-tion of Soil 2 was saturated with contaminants, and surplus contaminants may have been sorbed to the coarser fraction, which does not bind contaminants as tightly as the fine fraction, thus they can be more easily dissolved by solvents.

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The total removal efficiency from Soil 3, containing 40 % organic matter, was similar to that from the sandy soil ( Soil 2) with only 3.4 % organic matter. Given the high amount of organic content in Soil 3 the extractability from it was surprisingly high. However, Soil 3 contained large amounts of wood fibres (a previous study found the composition of the particulate matter in this soil to be similar to that of pine wood [143]), and wood sorption has been found to be weaker than expectations based on an organic carbon-based model (i.e. Koc) [175]. Indeed, Frankki et al. found that a high wood fibre content as a propor-tion of the total organic matter content (> 70 %) reduced the partitioning coef-ficient between the dissolved organic matter (DOM) and the particulate or-ganic content (POC) (log KPOC) by 0.5 units [143]. Hence, high amounts of wood fibres in a soil may reduce its adsorptive capacity and increase the mobil-ity of organic contaminants from such sites. This is a possible reason for the unexpectedly high extractability from the soil with high organic contents used in Paper III.

Table 8. Initial concentrations of PCDD/Fs in Soils 2, 3, and 4, residual concentra-tions in each soil and total removal efficiency after ten washing cycles at 60 °C

Soil number 1 2 3 4

Initial pg-TEQ/g 30 950 2300 8100

Estimated particle removal (%) 50 16 27 *

Total removal efficiency (%) 81 98 97 85

The lowest extractability was observed from the clay soil ( Soil 4, 3.7 % organic matter), in which graphite particles were also present. These results agree well with previous results since clay soils are generally not very suitable for soil washing due to their low permeability. Khodadoust et al. observed an extracta-bility of 29 % in one soil, compared to 19 % in a soil with higher clay contents [169]. The observed lower extractability from Soil 4 in Paper III may reflect the clay content in that soil. Furthermore, the graphite particles in Soil 4 also probably contributed to the lower extractability, since black carbon has ex-tremely high sorptivity for hydrophobic compounds like dioxins and furans [178, 179]. However particle removal from Soil 4 could not be estimated.

Hence, it can be concluded that the amount, rather than type, of organic con-tent affects the solvent-washing process under the applied conditions. Further, it will affect contaminant removal by particle removal more than the total re-moval efficiency.

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Influence of soil characteristics [Paper IV]

Although the amount of soil organic matter did not show any effect on the total extractability in Paper III, GAC adsorbed PCDD/Fs much more slowly from the organic soil extract than the sandy soil extract in Paper IV (Figure 10), and less carbon was needed to remove PCDD/Fs from the extract of the soil with low organic contents. The higher amount of carbon required for the high organic soil extract may have been due to co-extracted organic material competing for the adsorption sites on the GAC. The co-extracted material may hinder adsorption of the organic contaminants by blocking pores or via direct competition for sites between PCDD/F and DOM [178-180]. Furthermore, it has been shown that the contamination pattern is often quite complex at for-mer sawmill sites treated with CPs. Other organic chlorinated co-contaminants are often present in relatively high concentrations at these sites [183], and these co-contaminants were probably co-extracted during the washing process, along with the PCDD/Fs, and may thus also have competed for the GAC adsorp-tion sites.

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5.3 Feasibility of the remediation methods considered

As the results from this thesis indicate, all of the investigated methods for treating aged industrial soils described in Papers I-IV have limitations. How-ever, by adjusting and optimizing them, or combining them with other ap-proaches, it may be possible to limit or even avoid any drawbacks associated with the methods. Alternative approaches for the remediation methods con-sidered in Papers I-IV will therefore be discussed in this section.

Chemical oxidation To summarize some of the results from Papers I and II, the investigated chemical oxidation methods are least efficient for aged, fine-structured soils that are predominantly contaminated with HMW PAHs, and high organic con-tents will have a more negative effect on the degradability of LMW PAHs than of HMW PAHs. To overcome these limitations, certain measures must be taken. For instance, as already mentioned, using excess oxidant or agents that enhance the solubility of the contaminants may increase their availability, and thus the degradability of strongly sorbed compounds. The use of multiple oxi-dation cycles may also increase the degradation efficiency. For instance, Ari-enzo [182] found that TNT oxidation was more effective when Fenton’s re-agent was added in eight increments at 4 h intervals, instead of in a single or double dose. In Paper I the degradation efficiency after applying a second addition of hydrogen peroxide was investigated, but no increase in degradabil-ity was observed. However, considering the results presented by Arienzo [182], it seems reasonable to believe that the peroxide volume that was used as a sin-gle dose in Paper I may have been more effective if it had been split into sev-eral smaller doses.

Bioremediation methods are not generally as efficient as chemical oxidation methods for highly hydrophobic compounds, but chemical oxidation methods are unfortunately more expensive than bioremediation methods. However, by combining chemical remediation methods with biological treatments of aged, PAH-contaminated soils remediation can most likely be optimized with respect to both efficiency and costs. For this reason, the feasibility of chemically pre-treating PAH-contaminated soils in order to improve bioremediation has been extensively investigated [74, 75, 103]. In a study by Kulik et al., combinations of bioremediation and Fenton-like or ozone treatments provided greater total degradation efficiency of PAHs in creosote-contaminated sand and peat than any of the treatments used alone[102]. Nam et al. have reported similar obser-vations [53]. Similarly, Haapea and Tuhkanen achieved almost 90 % removal of

53

PAHs from soil by combining soil washing with ozonation and biodegradation, which was not possible using any of the methods alone [100]. Another major benefit of combining techniques is that it can reduce requirements for chemi-cals. Haapea and Tuhkanen were able to reduce the total ozone dose required by 50-70 % by prewashing the soil. The biological post-treatment also lowered the total ozone requirement [100].

Physical treatment The solvent-washing process can be adjusted according to the type of con-taminants present and/or soil characteristics, in order to reduce the costs and/or optimize the removal efficiency. As mentioned in section 3, solvent-washing is more suitable for coarse soils than for fine material. However, by applying particle size separation prior to treatment, the volume of material that needs to be treated can be dramatically reduced, or each fraction can be treated separately. In addition, the extraction mode can be altered. In Paper III the experiments were conducted as batch-wise extractions, but counter-current washing could significantly reduce the required solvent volumes, while proba-bly maintaining or improving the removal efficiencies For instance, a three-stage cross-current (or batch-wise) washing uses three times as much solvent as a counter-current process with similar removal efficiencies [169]. By using a simple water rinse stage after two counter-current wash cycles, costs could be further reduced since it would eliminate the need for a third washing cycle us-ing solvent [70]. Further, careful selection of washing solvent enables the treatment of different contaminants, and simultaneous removal of both metals and organic pollutants has been shown to be possible [58, 59].

After the contaminants are transferred to the solvent, there are numerous post-treatment options to choose from. In Paper IV the potential utility of adsorp-tion to granular activated carbon was investigated. It was found that the inves-tigated carbon did adsorb the targeted contaminants. However, considering the high hydrophobicity of PCDD/Fs, surprisingly large amounts of carbon were required to achieve high adsorption efficiency. This is a drawback since it would result in high costs for regenerating the GAC or high disposal costs of carbon filters. However, the viability of using GAC to remove PCDD/Fs from ethanol-based washing fluids should not yet be ruled out. Further, methods for minimizing the GAC volume could be applied. For instance, since NOM is known to adsorb to activated carbon, coextracted organic material in the wash-ing process may pose problems by blocking the pores of the filter and thus reducing the efficiency of the contaminant removal. To avoid clogging the filter and/or reducing the total organic load on the filter, the natural organic matter could be efficiently removed by using a pre-filter or either coagulating

54

or flocculating NOM prior to GAC filtration. Such techniques have already been developed and are being applied at full scale in drinking water purification processes [183].

Pore blockage can also be reduced by increasing the abundance of pores of the size range where NOM adsorbs [180]. The source material used for preparing GAC has been shown to have significant effects on the pore structure, surface texture, resistance to fragmentation, and adsorption capacity [184]. Hence, careful selection of an appropriate type of granular activated carbon for effi-cient PCDD/F removal is clearly important.

Reactivation of the carbon granules in the filter prolongs the filter’s lifetime. Matilainen et al. noted that adsorption initially increased after thermal regenera-tions of GAC granules [183]. However, within a few months the GAC removal declined to the same level as before the regeneration process. This may have been because the regeneration process changed the textural properties of the GAC, and consequently its adsorption preferences [183, 185]. Nevertheless, overall it can be concluded that GAC adsorption of dioxins is a realistic post-treatment step for solvent-washing liquids, provided that the type of activated carbon used is carefully selected, and appropriate regeneration processes for the GAC are developed and applied.

An alternative approach for GAC applications is to apply simultaneous adsorp-tion and contaminant degradation processes. This has been done, for instance, by combining an adsorption step with a biological treatment step [186, 187, 188]. Notably, in a pilot study by Khodadoust residual pentachlorophenol and solvent were treated sequentially in anaerobic and aerobic granular-activated carbon fluidized-bed reactors, after solvent recovery, and complete mineraliza-tion was achieved [70]. Liu et al. proposed a simultaneous degradation and re-generation process involving use of GAC adsorption/microwave (MW) regen-eration for treatment of a simulated PCB-containing soil-washing solution [185]. In addition, GAC impregnated with Fe/Pd bimetallic nanoparticles has been recently synthesised for dechlorination of PCBs [189, 190]. Although the results look promising, these recent GAC applications are in their infancy, and the costs of the reactive activated carbon need to be reduced before they can used in large-scale applications.

Other possible post-treatments of the washing fluids include distillation and UV irradiation. Distillation efficiently purifies and recovers solvents after wash-ing processes [56, 70], reducing the solvent volume and leaving a residue with high concentrations of the contaminants. The residue can be subsequently discharged at lower costs than the original contaminated volume, or subjected to a second post-treatment, possibly UV irradiation, since it has been proposed

55

by numerous authors to be a potential degradation method for dissolved PCDD/Fs [53, 55, 57, 63, 64]. However, careful solvent selection is essential, since the degradation efficiency depends on the solvent in which contaminants are dissolved. For example, Isosaari et al. found their degradability to be higher in olive oil than in palm oil or common organic solvents, such as toluene and hexane [64]. Therefore, when considering UV degradation, the solvent selec-tion should be based not only on its ability to dissolve the target contaminants, but also its transparency and potential capacity to consume hydroxyl radicals.

Finally, physical separation methods can, naturally, also be combined with chemical or biological degradation methods. As the contaminants are dissolved they become chemically available, which facilitates efficient degradation. Lee and Hosomi applied multiple ethanol washings of a soil [56], followed by distil-lation and Fenton treatment of the ethanol concentrate, obtaining removal efficiencies exceeding 99 % [168]. In another study of combined treatments Haapea and Tuhkanen examined the efficacy of successively applying several remediation methods (soil washing, ozonation and biological treatment) as an integrated treatment for PAH-contaminated soil. The consumption of ozone was 5-10 fold lower than without prewashing, and by combining the methods higher removal efficiency was achieved than when any of the methods were used alone [100].

Based on the results presented in Papers I-IV and previous investigations, it can be concluded that there are major advantages in combining methods. No-tably, it can reduce the adverse effects of contaminant sequestration and soil characteristics on the treatment results. Furthermore, combining techniques can make satisfactory treatment of a contaminated site more cost-effective.

Hence, I consider the methods investigated in this work to be highly feasible, provided that sufficient information about the site to be treated is acquired, and the methods to be applied are optimized prior to their application.

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6. CONCLUDING REMARKS AND FUTURE PERSPECTIVES

Although extensive research has been carried out to elucidate the factors influ-encing the sequestration of organic contaminants in soils, the processed in-volved are not yet fully understood and there are several important gaps in our knowledge.

Many of the parameters influencing sequestration and chemical degradability are probably the same as those influencing bioavailability. However, chemical oxidation and dissolution by organic solvents are usually faster processes than biodegradation and natural leaching processes. Consequently, chemical avail-ability and bioavailability are probably influenced by the same parameters, but to different extents.

Fenton’s reagent and ozone treatments are well-investigated chemical oxida-tion methods that have been found to be promising as either sole techniques or in combination with other techniques. The results in this thesis do not con-tradict earlier findings. However, they provide additional information regarding the effects of sequestration on chemical degradability. In Paper I clear differ-ences on the degradability of PAHs with chemical oxidation methods were observed between soils in which the contaminants had been aged to different extents. This finding highlights the great need to investigate the effects of soil properties on the sequestration and chemical degradation of organic contami-nants in aged soils. In Paper II the interactions between several factors were extensively investigated, using multivariate analysis, which identified interactive effects between individual physico-chemical properties of PAHs and soil pa-rameters that are known to influence their sequestration in aged soils. The re-sults in Paper II showed that the soil characteristics play important roles, but the physico-chemical properties of the contaminants must also be taken into consideration. Thus, it would be desirable to develop fast, reliable methods for determining soil parameters and the extent of sequestration of contaminants with certain properties. This could provide valuable indications of suitable remediation methods. For example, among soils predominantly contaminated with LMW PAHs, degradability is likely to be lower in soils with high organic contents. However, for soils contaminated with HMW PAHs, the oxidation

58

efficiency is more or less independent of the amount of organic matter. For such soils, enhancement of the solubility of the contaminants by using desorb-ing agents or aggressive oxidation conditions may be essential to degrade these compounds.

Solvent-washing is a relatively new treatment for dioxin-contaminated soils, which was shown, in Paper III, to have high removal efficiencies. Although the soil-washing technique could already be used in water-based applications, it has to be modified for use with solvents. Furthermore, it was shown that the amount of organic matter did not always affect the total removal efficiency during the solvent-washing described in Paper III, although the type of or-ganic matter did. Soil characteristics seem to have stronger effects on the mechanism responsible for removal of the contaminants (solvent dissolution versus particle removal), than on the amounts removed. Investigation of a pos-sible post-treatment of the washing fluid showed that the amount of organic matter in the soil substantially affected the regeneration of the solvent using granular activated carbon. Hence, it is important to consider the whole reme-diation chain, and perhaps develop several options for regenerating the solvent that can be applied in treatments of different soil types. Alternatively, the pre-treatment prior to GAC adsorption should be further developed, to minimize the effect of the organic matter for instance.

The method for regenerating the solvent described in Paper IV is still under development, and further modification of the process by pre-treatment of the solvent is most likely needed. Based on the results presented in Paper IV, in combination with previous results and knowledge gained regarding GAC ad-sorption of hydrophobic compounds, the technique looks promising. How-ever, an efficient post-treatment technique for removing water from the sol-vent is still required.

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The following proposals for future studies have emerged in the course of the studies underlying this thesis:

� In the work underlying this thesis, formation of degradation products was not investigated. However, since this is of great environmental concern, it would be interesting to continue the work by investigating whether formation of degradation products is affected by soils differ-ent characteristics, and if the formation of oxy-PAHs correlates with the degradation efficiency of their parent compounds.

� Since SOM is very heterogenic, further characterisation of soil organic matter and investigation of the impact of different SOM constituents on the chemical degradability would be of interest.

� Since the organic content in the soil negatively affected the PCDD/F adsorptivity of the GAC used to regenerate the washing solvent, the liquid purification process should be optimized by introducing an effi-cient NOM-removal process, and this process should be tested with regards to its compatibility with the washing liquid. The carbon filter also needs to be optimised in terms of its dimensions and structural properties.

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7. ACKNOWLEDGEMENTS

This work was partly performed within the Swedish national research program” Soil Remediation in a Cold Climate” (COLDREM), within financial support from the Foundation for Strategic Environmental Research (MISTRA) (Study I & II) and partly with support from the North Sweden Remediation Center (MCN), financed by the EU Structural Funds and New Objectives 1.(Study III & IV). I am very grateful for all the financial support that I recieved during my work.

Ett tack till Vägverket och inte minst Birger Sundström för ett trevligt samarbete och en osviklig (och mycket smittsam) entusiasm.

Tiden går fort när man har roligt säger man… Jag minns att jag började min doktorandtjänst med inställningen ”fyra år går kvickt”. Efter ”fyra” år med forskning måste jag dock säga att detta har varit de längsta fyra åren i mitt liv! Men då har det å andra sidan hunnit förflyta +8 år sen dess. Med tanke på det så gäller nog uttrycket fortfarande. För roligt har jag faktiskt haft, tack vare de runt mig som underlättat min tillvaro;

Jag har ”avverkat” ett flertal handledare under dessa år; Först ut; Bert van Bavel, du hann ge ett viktigt första intryck om att forskningsvärlden kan vara både seriös och rolig på samma gång. Mats Tysklind ryckte sen ut som huvudhandledare och Peter Haglund och Staffan Lundstedt tog sig an biträdande handledarskapet, alla med bra-vur. Ni har varit ett bra team som har stöttat, pushat samt gett en ovärderlig hjälp främst på det skriftliga planet. Trots mina infall med byte av jobb både en och två gånger har ni visat ett beundransvärt tålamod. Tack!

Det var ett tag sen jag jobbade hos er på ”gamla Miljökemi” (vad ni heter nu är för mig en gåta!). Jag minns många trevliga, stolliga och intressanta diskussioner i fika-rummet på plan 7. För dessa och alla glada skratt säger jag tack; till er som är kvar och även ni som rört på er.

Ett speciellt tack till; Staffan för att du delat med dig om dina kunskaper i PAH analy-ser och uppreningsmetoder, Rolf, för all hjälp med dioxinanalyser, Ylva, min pärla, likt en hamster samlade du på dig jordprover som du frikostigt delade med dig av tillsam-mans med massor av bakgrundskunskaper och trevliga (ej jobbrelaterade) samtalsäm-nen.; Eva K, jag kan fortfarande sakna våra många samtal, men vi får se till ta igen det på annat sätt; Lena Burström, har förmodligen kammat hem flest tack på f.d. Miljö-kemi under de år jag var där, med all rätta. Du var navet för Miljökemi, hur det går för dem utan dig det undrar jag. Framför allt tack för ditt trevliga sällskap, men även stöd och hjälp med allt mellan himmel och jord; Barbro, när man vänder sig till dig med en

62

fråga får man alltid ett rakt, ärligt och förnuftigt svar, med kvinnlig touch! Därför var du viktig för mig på miljökemi; Till sist, Stina, tusen tack för att jag fick ha dig som bollplank under min mest intensiva skrivningsperiod med allt vad det innebar; hjälp med att skicka mail i rätt riktning, lån av arbetsplats, dator osv. Men framför allt, tack för att du sprider glädje runt dig.

Ett stort tack till gamla och nuvarande arbetskamrater på ”Styret” och FMV som har haft tålamod med mig och detta hobbyprojekt…eller visst har ni haft det?

Anna-Karin, Helen och Kickan, vi har känt varandra i 30 år nu. Jag är tacksam för att ni finns i mitt liv. I mångt och mycket har er vänskap format mig till den jag är idag. Så se där vad ni har orsakat!

Plugga på universitetet var en spännande tid. Varför kändes den så tung då? Nu minns jag bara fondueaftnar med Jenny som bartender, (tur att du valde ett annat yrke!), Tina som hjular hem (har du tränat mer på det?) och Marias nudlar som serverades med tegelpannor och hårstrån i Vietnam (har du ätit Mi xau Rau på sistone?) Tack för allt kul vi haft och kommer att ha.

Andra härliga vänner som jag vill tacka för att de vill vara just det är; Jenny B, kära kusin, vår vänskap har tagit många skepnader genom åren. Nu ägnar vi oss åt det som våra mammor gjort, delar orostankar och glädjeämnen över barnen, låt oss dock för-söka fokusera mest på det senare. Annika, min riviga mammakollega jag har alltid kul dig, Nina, alla ord är överflödiga när man beskriver dig. You’re one in a million, bättre än den finaste kristall!

Mamma, pappa; även om ni ofta frågat vad jag ”egentligen” gör, så har ni stöttat mig i mina val och det är jag tacksam för. Bror, du har stöttat mig men också härdat mig genom åren med diverse påhitt och fusk i allehanda spel. Tack vare dig vet jag att världen kan vara falsk men ganska kul trots det.

Jag vill också rikta ett tack till (ännu) icke ingift släkt. Speciellt tack till Paula och Hå-kan, framför allt för att ni har en så fin son! Och Mariana, vår alldeles egna proffs-barnvakt, killarna är tokiga i dig. Vi med!

Nu till ”das kärn des pudels” som en viss gotlänning skulle ha sagt…

Jag vill tack alla er, ”Grabbarna Grahn” för att ni har ställt upp på så mycket på mig. Jag hoppas jag får möjlighet att göra detsamma för er en vacker dag.

Hans, du har dragit ett stort lass med husrenovering, veckopendling och nu mitt av-handlingsarbete. Men nu är detta till ände och nu kan vi bara njuta av vår tid tillsam-mans. Jag älskar dig, det vet du! Albin, jag vill att du ska veta att jag är fantastiskt stolt över dig och älskar dig. Småkillarna kan inte ha bättre storebror och förebild. Eskil och Vilgot, äntligen är mamma klar med ”himla boken”. Nu får jag vara hemma med er och vara mer fokuserad på ”här och nu”. Jag älskar er oändligt. Till sist och definitivt minst; Lille Melker, när jag först började på denna kappa låg du och skvalpa-de i min mage. Du gav mig kraft och motivation att sätta igång med att slutföra detta. Jag älskar dig som en tok!

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[1] Weber R., Gaus C., Tysklind M., Johnston P., Forter M., Hollert H., Heinisch E.,

Holoubek I., Lloyd-Smith M., Masunaga S., Moccarelli P., Santillo D., Seike N., Symons R., Machado Torres J.P., Verta M., Varbelow G., Vijgen J., Watson A., Costner P., Woelz J., Wycisk P., Zennegg M., Dioxin- and POP-contaminated sites-contemporary and future relevance and challenges, Environ. Sci. Pollut. Res. 15 (2008) 363-393.

[2] Giftfri miljö, Miljökvalitetsmål 12, Kemikalieinspektionens rapport, (1999) [Non-toxic environment, Swedish Chemicals Agency report] Norstedts Tryckeri AB Stockholm, Sweden, ISBN 91-7932-041-4 Available at: www.kemi.se/upload/pdf/Trycksaker/Broschyrer/GiftfriMiljo1999.pdf

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APPENDIX

Personal thoughts on remediation in Sweden.

There is a great, urgent need for environmentally safe, cost-effective methods for remediating contaminated sites. Many possible remediation methods have been proposed, presented and discussed in the research literature, but far fewer have been sufficiently developed for routine, commercial, large-scale applica-tion. In the long term, the key factors determining whether or not a remedia-tion method will be scaled up from laboratory tests to full-scale application are its performance, cost and the perceived need for it, balanced against budgetary restrictions. The relative strength of these factors will vary widely between dif-ferent countries. For instance, in many places in the Third World there may be an extreme need for remediation, but no funds for remedial treatments. In various industrialised countries, where funds are more readily available, reme-diation programs are being intensively developed. In Sweden, however, treat-ment methods and capacities have not been as fully developed as in many other industrialised countries.

In my opinion, several reasons can be identified for the relative lack of interest in developing full-scale techniques in Sweden, one important being that the costs of land disposal of contaminated soil volumes are too low in Sweden. Hence, there is little incentive to invest large sums of money in developing remediation techniques (or using existing ones). Instead, large volumes of soil are affordably deposited, in many cases, rather than using techniques that elim-inate the contaminants, or convert them into safe forms. In fact landfill dispos-al is the main remedial activity in Sweden and during 2000 and 2004, 550 000 tons of contaminated material was disposed on a yearly basisi.ii Another factor is the demand for unrealistically high remediation efficiencies. Leaving pollu-tants behind after treatment is not consistent with political objectives, notably the ‘1000-year perspective’ stipulated by the Swedish EPA (S-EPA), even if the contaminants are virtually unavailable and have very low leachability. This ex-cludes many remediation methods and/or makes remediation extremely costly, which again favours deposition, or in some cases incineration.

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What can be done to overcome this problem?

Most importantly, the costs of deposition should be increased to a sufficiently high level to promote the use and development of destructive remediation methods. In Sweden, a decline in deposition has been observed after the intro-duction of taxes for waste disposal. Hence the wanted effect was achieved1. However further reduction of waste disposal is desirable. Provided that cost-effective, environmentally safe methods can be implemented, and realistic pre-dictions of the post-treatment availability of contaminants can be acquired, it should be possible to leave acceptable fractions of contaminants in the ground after careful monitoring of treated sites. Suér et al. reviewed remediation meth-ods from a Life Cycle Assessment (LCA) perspective recentlyiii, and concluded that site remediation will have both positive and negative environmental ef-fects, regardless of the technique applied. Obviously, the removal or degrada-tion of contaminants is a positive impact, whereas global warming through the energy consumed in transporting or heating soil, pumping, resource depletion, smog formation and noise are examples of negative effects. Including remain-ing contaminants in soil after treatment in the LCA can be important for the selection of technique. The amounts, availability and potential effects of con-taminants left in the soil after treatment should be considered in the LCA, and when selecting techniques to apply. Together with time and economic data, LCA can provide a sound basis for selecting appropriate remediation tech-niques in which the total environmental impacts are accounted for and made more readily apparent. LCA can also be used before and after site remediation, thus providing more detailed information and understanding that may facilitate the selection of optimal techniques for future remediation programs.

The S-EPA provides grants for investigations and remediation of sites in Swe-den in cases where no one is found to be responsible for contamination. If such sites are to be remediated, contractors should be chosen in compliance with the law on public procurement (LPP). According to S-EPA policies, pri-ority should then be given to destructive treatment solutions (when possible) and thereafter to concentrating technologies that reduce the volume of con-taminated material. Land disposal, immobilization or encapsulations are the least desirable solutions. Finally, prior to the selection of remediation tech-niques, a cost/benefit analysis should be performed, in which technical, eco-nomic and environmental factors are considered. If long-term consequences are uncertain, environmentally safe technologies should be selected, even if the costs are higher than those of other optionsiv. The LPP gives the S-EPA a po-tentially very effective tool, to limit the ‘dig-and-dump-mentality’ and open up possibilities to choose treatment technologies that are more consistent with the

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government’s stated policies by drafting appropriate specifications and contract terms. I question the extent to which this is done.

In order to approach the ‘Non-toxic environment’ objective, a large number of sites with contaminated soils need to be remediated. When doing so, we clearly do not want to cause new or additional problems. Therefore, there is a ten-dency to choose extremely “safe methods”, which essentially excludes all op-tions except incineration. However as already mentioned this treatment is very expensive and can only be realistically applied to the most highly prioritised sites. By increasing research into risk estimations for pollutants remaining in remediated soils, we may gain the confidence to continue leaving significant amounts of pollutants behind. Further, the reuse of treated soils should be encouraged. Today soil masses treated off site is not being backfilled at the site. To exclude risks for transportation of remaining contaminants from the con-taminated site or treated soil masses, leaching tests are recommended. How-ever, this research should be performed on site and not in a laboratory. The same approach should be applied to full-scale remediation trials, which should be actively encouraged and supported at test sites with relevant contamination patterns that are unlikely to pose great risks since they are located in remote areas with low population densities, and exposure to humans and wildlife is likely to be minimal. This has been done in the USA within the superfund pro-gramsv. In my opinion, this approach may provide us with cost efficient meth-ods that could be reliably applied to larger, more complex risk areas. By doing so we would also hopefully be able to afford to remediate larger numbers of contaminated sites and come closer to the goal of “A Non-toxic environment”.

i SOU, En BRASkatt- beskattning av avfall som deponerats, SOU 2005:64, Statensoffentliga utredningar, Regerings-kansliet, Stockholm 2005.

ii Van Hees P., Elgh-Dalgren K., Engwall M., von Kronhelm T., Re-cykling of remediated soil in Sweden: An envi-ronmental advantage? Resour. conserv. recycl. 52 (2008) 1349-1361.

iii Suér P., Nilsson-Påledal S., Norrman J., LCA for site remediation: A literature review, Soil & sediment contamina-tion. 13 (2004) 415-425.

iv Swedish Environmental Protection Agency, Guidance for the remediation of wood impregnation plants, (1999) Report 4963 [Naturvårdsverket, Vägledning för efterbehandling vid träskyddsanläggningar, (1999) Rapport 4963 Swedish Environmental Protection Agency. ISBN 91-120-4963-1.

v United States Environmental Protection Agency, Field applications of In Situ remediation technologies, Chemical Oxidations, Contract 68- W6-0014.