This report contains the collective views of an international group of experts and does not necessarily represent the decisions or the stated policy of the United Nations Environment Programme, the International Labour Organization, or the World Health Organization.

Concise International Chemical Assessment Document 60

CHLOROBENZENES OTHER THAN HEXACHLOROBENZENE:
ENVIRONMENTAL ASPECTS

First draft prepared by H.M. Malcolm, P.D. Howe, and S. Dobson, Centre for Ecology

and Hydrology, Monks Wood, United Kingdom

Published under the joint sponsorship of the United Nations Environment Programme, the International Labour Organization, and the World Health Organization, and produced within the framework of the Inter-Organization Programme for the Sound Management of Chemicals.

World Health Organization

Geneva, 2004

The International Programme on Chemical Safety (IPCS), established in 1980, is a joint venture of the United Nations Environment Programme (UNEP), the International Labour Organization (ILO), and the World Health Organization (WHO). The overall objectives of the IPCS are to establish the scientific basis for assessment of the risk to human health and the environment from exposure to chemicals, through international peer review processes, as a prerequisite for the promotion of chemical safety, and to provide technical assistance in strengthening national capacities for the sound management of chemicals.

The Inter-Organization Programme for the Sound Management of Chemicals (IOMC) was established in 1995 by UNEP, ILO, the Food and Agriculture Organization of the United Nations, WHO, the United Nations Industrial Development Organization, the United Nations Institute for Training and Research, and the Organisation for Economic Co-operation and Development (Participating Organizations), following recommendations made by the 1992 UN Conference on Environment and Development to strengthen cooperation and increase coordination in the field of chemical safety. The purpose of the IOMC is to promote coordination of the policies and activities pursued by the Participating Organizations, jointly or separately, to achieve the sound management of chemicals in relation to human health and the environment.

WHO Library Cataloguing-in-Publication Data

Chlorobenzenes other than hexachlorobenzene : environmental aspects.

(Concise international chemical assessment document ; 60)

1.Chlorobenzenes 2.Risk assessment 3.Environmental exposure

I.International Programme on Chemical Safety II.Series

ISBN 92 4 153060 X         (LC/NLM Classification: QV 633)

ISSN 1020-6167

©World Health Organization 2004

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Technically and linguistically edited by Marla Sheffer, Ottawa, Canada, and printed by Wissenchaftliche Verlagsgesellschaft mbH, Stuttgart, Germany

FOREWORD

Concise International Chemical Assessment Documents (CICADs) are the latest in a family of publications from the International Programme on Chemical Safety (IPCS) — a cooperative programme of the World Health Organization (WHO), the International Labour Organization (ILO), and the United Nations Environment Programme (UNEP). CICADs join the Environmental Health Criteria documents (EHCs) as authoritative documents on the risk assessment of chemicals.

International Chemical Safety Cards on the relevant chemical(s) are attached at the end of the CICAD, to provide the reader with concise information on the protection of human health and on emergency action. They are produced in a separate peer-reviewed procedure at IPCS. They may be complemented by information from IPCS Poison Information Monographs (PIM), similarly produced separately from the CICAD process.

CICADs are concise documents that provide summaries of the relevant scientific information concerning the potential effects of chemicals upon human health and/or the environment. They are usually based on selected national or regional evaluation documents or on existing EHCs. Before acceptance for publication as CICADs by IPCS, these documents undergo extensive peer review by internationally selected experts to ensure their completeness, accuracy in the way in which the original data are represented, and the validity of the conclusions drawn.

The primary objective of CICADs is characterization of hazard and dose–response from exposure to a chemical. CICADs are not a summary of all available data on a particular chemical; rather, they include only that information considered critical for characterization of the risk posed by the chemical. The critical studies are, however, presented in sufficient detail to support the conclusions drawn. For additional information, the reader should consult the identified source documents upon which the CICAD has been based.

Risks to human health and the environment will vary considerably depending upon the type and extent of exposure. Responsible authorities are strongly encouraged to characterize risk on the basis of locally measured or predicted exposure scenarios. To assist the reader, examples of exposure estimation and risk characterization are provided in CICADs, whenever possible. These examples cannot be considered as representing all possible exposure situations, but are provided as guidance only. The reader is referred to EHC 170.¹

While every effort is made to ensure that CICADs represent the current status of knowledge, new information is being developed constantly. Unless otherwise stated, CICADs are based on a search of the scientific literature to the date shown in the executive summary. In the event that a reader becomes aware of new information that would change the conclusions drawn in a CICAD, the reader is requested to contact IPCS to inform it of the new information.

Procedures

The flow chart on page 2 shows the procedures followed to produce a CICAD. These procedures are designed to take advantage of the expertise that exists around the world — expertise that is required to produce the high-quality evaluations of toxicological, exposure, and other data that are necessary for assessing risks to human health and/or the environment. The IPCS Risk Assessment Steering Group advises the Coordinator, IPCS, on the selection of chemicals for an IPCS risk assessment based on the following criteria:

• there is the probability of exposure; and/or
• there is significant toxicity/ecotoxicity.

Thus, it is typical of a priority chemical that

• it is of transboundary concern;
• it is of concern to a range of countries (developed, developing, and those with economies in transition) for possible risk management;
• there is significant international trade;
• it has high production volume;
• it has dispersive use.

The Steering Group will also advise IPCS on the appropriate form of the document (i.e., a standard CICAD or a de novo CICAD) and which institution bears the responsibility of the document production, as well as on the type and extent of the international peer review.

The first draft is usually based on an existing national, regional, or international review. When no appropriate source document is available, a CICAD may be produced de novo. Authors of the first draft are usually, but not necessarily, from the institution that developed the original review. A standard outline has been developed to encourage consistency in form. The first draft undergoes primary review by IPCS to ensure that it meets the specified criteria for CICADs.

The second stage involves international peer review by scientists known for their particular expertise and by scientists selected from an international roster compiled by IPCS through recommendations from IPCS national Contact Points and from IPCS Participating Institutions. Adequate time is allowed for the selected experts to undertake a thorough review. Authors are required to take reviewers’ comments into account and revise their draft, if necessary. The resulting second draft is submitted to a Final Review Board together with the reviewers’ comments. At any stage in the international review process, a consultative group may be necessary to address specific areas of the science. When a CICAD is prepared de novo, a consultative group is normally convened.

The CICAD Final Review Board has several important functions:

• to ensure that each CICAD has been subjected to an appropriate and thorough peer review;
• to verify that the peer reviewers’ comments have been addressed appropriately;
• to provide guidance to those responsible for the preparation of CICADs on how to resolve any remaining issues if, in the opinion of the Board, the author has not adequately addressed all comments of the reviewers; and
• to approve CICADs as international assessments.

Board members serve in their personal capacity, not as representatives of any organization, government, or industry. They are selected because of their expertise in human and environmental toxicology or because of their experience in the regulation of chemicals. Boards are chosen according to the range of expertise required for a meeting and the need for balanced geographic representation.

Board members, authors, reviewers, consultants, and advisers who participate in the preparation of a CICAD are required to declare any real or potential conflict of interest in relation to the subjects under discussion at any stage of the process. Representatives of nongovernmental organizations may be invited to observe the proceedings of the Final Review Board. Observers may participate in Board discussions only at the invitation of the Chairperson, and they may not participate in the final decision-making process.

This CICAD on chlorobenzenes other than hexachlorobenzene (environmental aspects) is an update of Environmental Health Criteria (EHC) 128, Chlorobenzenes other than hexachlorobenzene (IPCS, 1991a). Information on the fate and levels of chlorobenzenes was also obtained from Agency for Toxic Substances and Disease Registry reports on chlorobenzene (ATSDR, 1990) and 1,4-dichlorobenzene (ATSDR, 1998). A further literature search was performed up to December 2002 to identify any additional information published since these reviews were completed. Information on the peer review of the source document is presented in Appendix 1. Information on the peer review of this CICAD is presented in Appendix 2. This CICAD was approved as an international assessment at a meeting of the Final Review Board, held in Varna, Bulgaria, on 8–11 September 2003. Participants at the Final Review Board meeting are listed in Appendix 3. The International Chemical Safety Cards for a number of different chlorobenzenes (ICSC 0037, 0344, 0531, 0642, 0676, 1049, 1066, 1095, 1222), produced by the International Programme on Chemical Safety (IPCS, 2000, 2003a–h), have also been reproduced in this document. This CICAD concentrates on environmental aspects because there have been no significant changes to the human health assessment since publication of the EHC (IPCS, 1991a).

Chlorinated benzenes are a group of cyclic aromatic compounds in which one or more hydrogen atoms of the benzene ring have been replaced by a chlorine atom. Chlorobenzenes are used mainly as intermediates in the synthesis of pesticides and other chemicals. 1,4-Dichlorobenzene (1,4-DCB) is used in space deodorants and as a moth repellent. The higher chlorinated benzenes (trichlorobenzenes, 1,2,3,4-tetrachlorobenzene [1,2,3,4-TeCB], and pentachlorobenzene [PeCB]) have been used as components of dielectric fluids.

Natural sources of chlorobenzenes in the environment have not been identified. Chlorobenzenes are released to the environment during manufacture or use as intermediates in the production of other chemicals. They will also be released during the disposal of chlorobenzene products, such as from incinerators and hazardous waste sites. Monochlorobenzene (MCB) is released directly to the environment due to its use as a pesticide carrier. Chlorobenzenes used as deodorizers, fumigants, degreasers, insecticides, herbicides, and defoliants will also be released to the environment as a direct result of their application.

Their physicochemical properties suggest that chlorobenzenes released to the environment are likely to volatilize to the atmosphere. Removal of chlorobenzenes from the atmosphere will occur primarily via reactions with hydroxyl radicals to produce nitrochlorobenzene, chlorophenol, and aliphatic dicarbonyl products, which are further removed by photolysis or reaction with hydroxyl radicals. Chlorobenzenes released into the aquatic environment will be redistributed preferentially to the air and to sediment (particularly organically rich sediments). Chlorobenzenes in aqueous solutions could, in theory, undergo photochemical reductive dechlorination, although studies have been performed only under artificial conditions that were not representative of temperate regions. The most important factor affecting the behaviour and fate of chlorobenzenes in soil is sorption. Adsorption–desorption processes in soil affect the rate of volatilization and leaching and the availability of chemicals to microbial and chemical degradation or uptake by plants or other organisms.

Chlorobenzenes in various substrates, including soil, sediment, and sewage sludge, may be degraded by microorganisms. The major mechanism of aerobic degradation is via oxidative dechlorination, leading to the formation of hydroxylated aromatic compounds (mainly catechols), which undergo ring fission and subsequent mineralization to carbon dioxide and water. The less chlorinated benzenes are more readily degraded than the higher chlorinated ones.

The bioaccumulation of chlorobenzenes by aquatic organisms is determined by their relative water and lipid solubilities (thus reflecting the octanol/water partition coefficients) and the number of chlorine substitutions. Uptake from water increases with increasing chlorination and increasing temperature.

Concentrations of chlorobenzenes (MCB, dichlorobenzenes, and trichlorobenzenes) have been reported in ambient air, with mean concentrations in the order of 0.1 µg/m³ and maximum levels (at hazardous waste sites) of up to 100 µg/m³. Concentrations of chlorobenzenes in surface waters are generally in the ng/litre to µg/litre range, with maximum concentrations up to 0.2 mg/litre in areas close to industrial sources. Levels of chlorobenzenes in industrial wastewaters may be higher and vary according to the nature of the processes used. Chlorobenzene levels in uncontaminated soils are generally less than 0.4 mg/kg for dichlorobenzene congeners and less than 0.1 mg/kg for other chlorobenzene congeners. Levels of chlorobenzenes in sediments are generally in the ng/kg to µg/kg range, although levels in the mg/kg range have been reported in samples from industrial areas.

In general, aquatic toxicity increases with the degree of chlorination of the benzene ring. Seventy-two-hour EC₅₀s for green algae range from 5280 µg/litre for 1,3-DCB to 200 000 µg/litre for MCB; similarly, 48-h EC₅₀s for diatoms range from 8 to 235 000 µg/litre. For freshwater invertebrates, 48-h EC₅₀s range from 10 µg/litre for PeCB to >530 000 µg/litre for 1,2,4,5-TeCB. Ninety-six-hour LC₅₀s for fish range from 135 µg/litre for PeCB to 21 000 µg/litre for 1,2,4-trichlorobenzene (1,2,4-TCB). Chronic no-observed-effect concentrations (NOECs) for freshwater invertebrates range from 32 µg/litre for PeCB to 19 000 µg/litre for MCB; in fish, NOECs range from 18 µg/litre for PeCB to 8500 µg/litre for MCB.

Few data are available on the effects of chlorobenzenes on terrestrial systems. LC₅₀ values for plants grown hydroponically or in soil ranged from 0.028 to 9.3 mg/litre and from 1 to >1000 mg/kg soil, respectively. LC₅₀ values for the earthworms Eisenia andrei and Lumbricus rubellus ranged from 0.22 µmol/litre (pore water) for PeCB to 4281 µmol/litre for MCB.

The risk of chlorinated benzenes causing harm to aquatic organisms is low. Risk factors comparing chronic toxicity values with concentrations measured in the environment were generally below 1, with the exception of some compounds that had higher risk factors, with a maximum value of 200. The highest risk factors were derived using old data from point sources and are therefore unrepresentative of the whole environment, especially when the likelihood of evaporation is considered. There were inadequate data to perform a risk assessment for terrestrial species.

Chlorinated benzenes are a group of cyclic aromatic compounds in which one or more hydrogen atoms of the benzene ring have been replaced by a chlorine atom. The generic molecular formula is C₆H_6–nCl_n, where n = 1–6. There are 12 different chlorinated benzenes: monochlorobenzene (MCB), dichlorobenzene (DCB) (three isomers), trichlorobenzene (TCB) (three isomers), tetrachlorobenzene (TeCB) (three isomers), pentachlorobenzene (PeCB), and hexachlorobenzene. Hexachlorobenzene is reviewed in a separate EHC (IPCS, 1997) and is therefore not covered by this CICAD.

The identity of chlorobenzenes and their physical and chemical properties are presented in Table 1. MCB, 1,2-DCB, 1,3-DCB, and 1,2,4-TCB are colourless liquids, while all other congeners are white crystalline solids at room temperature. In general, the solubility of chlorobenzenes in water is low (decreasing with increasing chlorination), flammability is low, the octanol/water partition coefficients are moderate to high (increasing with increasing chlorination), and vapour pressures are low to moderate (decreasing with increasing chlorination) (IPCS, 1991a).

Table 1: Physicochemical properties of chlorobenzenes.^a

The analytical technique of choice for the determination of chlorobenzenes in environmental samples is gas chromatography (GC). However, the methods of collection and preparation of samples for GC analysis vary considerably, depending on the medium and the laboratory. Capillary columns with different stationary phases are frequently used to separate compounds. Detection occurs via the use of a flame ionization detector (FID), electron capture detector (ECD), or mass spectrometric (MS) detector (IPCS, 1991a).

Tenax-GC resins have commonly been used as adsorbents for the air sampling of chlorobenzenes (Krost et al., 1982; Pellizzari et al., 1982), although XAD resins have also been used (Langhorst & Nestrick, 1979). Air pollutants collected on Tenax-GC resins can be desorbed directly onto the GC column by heating the tube with sorbent. XAD resins can be extracted with solvents, an aliquot of which can then be injected into a GC. Detection limits in the 1970s ranged from 0.7 µg/m³ for MCB to 0.9 µg/m³ for PeCB (Langhorst & Nestrick, 1979); however, much lower detection limits have been achieved more recently using ECD (0.5 pg/m³ for PeCB to 1.8 pg/m³ for 1,2,4,5-TeCB) (Hermanson et al., 1997).

Solvent extraction is a simple and effective technique for recovering chlorobenzenes from water samples. Hexane, pentane, and a 1:1 mixture of cyclohexane and diethyl ether have been identified as suitable extraction solvents for these compounds (Oliver & Bothen, 1980; Piet et al., 1980; Otson & Williams, 1981; Meharg et al., 2000). Alternatively, preconcentration of the chlorobenzenes on organic resins, such as Chromosorb 102 and Tenax-GC, is also effective; detection limits using Chromosorb 102 were reported to range from 0.5 µg/litre for MCB to 0.01 ng/litre for PeCB (Oliver & Bothen, 1980; Pankow & Isabelle, 1982). The purge-and-trap method has also been used to concentrate the volatile halogenated benzenes before analysis using GC (Jungclaus et al., 1978; Pereira & Hughes, 1980; Otson & Williams, 1982; Huybrechts et al., 2000; Martinez et al., 2002). Detection limits of 0.1–0.2 µg/litre for MCB and various dichlorobenzene isomers were achieved using FID and Hall electrolyte conductivity detectors (Otson & Williams, 1982), 0.08 µg/litre for 1,2,4-TCB using ECD (Martinez et al., 2002), and 0.76–20 ng/litre for di- and trichlorobenzenes using MS (Huybrechts et al., 2000). More recently, alternative extraction techniques such as headspace solid-phase microextraction with GC-MS have achieved detection limits for individual chlorobenzene isomers ranging from 4 to 6 ng/litre (He et al., 2000); however, it should be noted that analytical techniques using simple solvent extraction and GC-MS can now attain detection limits ranging from 5 pg/litre for 1,2,3- and 1,3,5-TCB to 15 pg/litre for PeCB (Meharg et al., 2000).

The extraction of chlorobenzenes from aquatic sediments, sewage sludges, or soil can be achieved by solvent or Soxhlet extraction (Oliver & Bothen, 1982; Lopez-Avila et al., 1983; Onuska & Terry, 1985; Wang & Jones, 1991; Wang et al., 1992). Solvents commonly used are acetone and/or hexane. Other extraction methods, such as sonication, saponification, and supercritical fluid extraction, have been used to extract sediment-bound chlorobenzenes, but were found to be less efficient than Soxhlet extraction (Prytula & Pavlostathis, 1996). The extract is generally dried using sodium sulfate, followed by cleanup on a Florisil column before GC analysis with ECD, with detection limits of 1500 µg/kg for MCB and lower detection limits ranging from 1.5 µg/kg for dichlorobenzenes to 0.05 µg/kg for PeCB (Oliver & Bothen, 1982; Onuska & Terry, 1985; Wang & Jones, 1991; Wang et al., 1992). Alternatively, headspace solid-phase microextraction with GC–ion trap MS has been found to reproduce detection limits of 0.03–0.1 µg/kg for 1,2,3-TCB, 1,2,3,4-TeCB, and PeCB in soil (Santos et al., 1997).

For the detection of chlorobenzenes in biota samples, solvent or Soxhlet extraction with subsequent cleanup on Florisil columns and GC analysis with ECD have commonly been used (Lunde & Ofstad, 1976; Kuehl et al., 1980; Oliver & Bothen, 1982; Muir et al., 1992; Gebauer & Weseloh, 1993; Cobb et al., 1994; Jan et al., 1994; Wade et al., 1998). Detection limits of 1500 µg/kg for MCB and lower detection limits ranging from 5 µg/kg for dichlorobenzenes to 0.02 µg/kg for PeCB have been reported (Oliver & Bothen, 1982; Cobb et al., 1994). Vacuum extraction and the direct purge-and-trap method have also been used to quantify levels of MCB in fish tissue (Hiatt, 1981).

Natural sources of chlorobenzenes in the environment have not been identified. However, 1,2,3,4-TeCB has been identified in the oil of marsh grass, although it is not known whether this was formed naturally (Miles et al., 1973).

Chlorobenzenes are released to the environment from sites where they are either manufactured or used as intermediates in the production of other chemicals. They will also be released during the disposal of chlorobenzene products, such as from incinerators (IPCS, 1991a) and hazardous waste sites (ATSDR, 1998). Chlorobenzenes are a product of incomplete combustion and may therefore be released to the environment from waste incinerators. Chlorobenzenes may be formed from the metabolic breakdown of lindane in higher organisms and from its physical breakdown under extreme environmental conditions (IPCS, 1991b).

Releases of some chlorobenzene compounds to the environment in the USA in 2001, as recorded in the US Toxics Release Inventory (TRI), are listed in Table 2. These data do not form a comprehensive list, as only certain types of industrial facility are required to register in the TRI (ATSDR, 1998). There is a paucity of data on the quantity of chlorobenzenes released to the environment in other parts of the world, although some production and consumption data are available. Approximately 15 000 tonnes of 1,4-DCB were produced in and/or imported into the European Union in 1994 (EC, 2001). Total production of MCB, 1,2-DCB, and 1,4-DCB in Japan in 1998 was 26 351 tonnes (Chemical Daily Company, 1999), with 9073 tonnes imported in 1998 and 8310 tonnes imported in 1999 (Chemical Daily Company, 2000).

Table 2: Total releases of chlorobenzenes in the USA during 2001.^a

^a From US EPA (2003).

Some uses of chlorobenzenes, including uses as deodorizers, fumigants, degreasers, insecticides, herbicides, and defoliants, will result in direct releases to the environment.

MCB will be released directly to the environment due to its use as a pesticide carrier (Meek et al., 1994c). MCB is used as a solvent carrier for pesticides (29 000 kg per annum in Canada), in the manufacture of rubber polymers (20 000 kg per annum in Canada), and as a carrier for textile dyes (1000 kg per annum in Canada) (Mackay et al., 1996). Fifty per cent of the MCB used in Canada is released to the environment; 80% is emitted to the atmosphere, 10% to water, and 10% to soil, giving releases of 20 000, 2500, and 2500 tonnes, respectively, per year (Mackay et al., 1996). MCB is used in the production of phenol and nitrochlorobenzene (ortho and para isomers), in the formulation of herbicides, to produce additional chlorobenzenes, and as a solvent in the manufacture of adhesives, paints, resins, dyestuffs, and drugs (Grosjean, 1991). MCB is used in the manufacture of diphenyl oxide, phenylphenol, silicone resin, and other halogenated organics (ATSDR, 1990).

1,2-DCB is used primarily in the automotive and metal industries as a solvent for the removal of carbon and degreasing of metal parts (Meek et al., 1994a). 1,2-DCB is used in the synthesis of organic chemicals such as toluene diisocyanate (Grosjean, 1991).

1,4-DCB is used in air fresheners, urinal deodorants, and moth and bird repellents (Meek et al., 1994b; EC, 2001). All of these uses release 1,4-DCB to the environment, principally the atmosphere. 1,4-DCB is also used as an intermediate in the production of other chemicals, including polyphenylene sulfide resins (Grosjean, 1991) and 1,2,4-TCB (ATSDR, 1998). Minor uses of 1,4-DCB include its use in the control of tree-boring insects, ants, and blue mould in tobacco seedbeds (ATSDR, 1998).

Trichlorobenzenes, especially 1,2,4-TCB, are used as dye carriers, degreasing solvents, oil additives, and dielectric fluids and in the formulation of pesticides (Grosjean, 1991). The use of trichlorobenzenes is restricted to mainly 1,2,4-TCB, which is used as a chemical intermediate and an industrial solvent (Giddings et al., 1994c). 1,2,4-TCB was formerly used as a degreasing agent, in septic tanks, and in drain cleaners, wood preservatives, and abrasive formulations (EC, 2003).

Tetrachlorobenzenes and pentachlorobenzenes may be released to the environment from the spillage of dielectric fluids (Giddings et al., 1994a,b). 1,2,3,4-TeCB is used as a component in dielectric fluids (IPCS, 1991a). 1,2,4,5-TeCB is used as an intermediate in the manufacture of herbicides and defoliants. It is also used as an insecticide, as a moisture-resistant impregnant, in electrical insulation, and in packing protection (IPCS, 1991a). PeCB was formerly used in a pesticide to combat oyster drills (small snails that eat oysters). It has also been used as an intermediate (IPCS, 1991a).

Their physicochemical properties suggest that chlorobenzenes released to the environment are likely to be volatilized to the atmosphere. The Henry’s law constants measured for chlorobenzenes suggest that they are readily volatilized, especially from aquatic systems with long residence times, such as large lakes and oceans (Ten Hulscher et al., 1992). However, chlorobenzenes released to water may also be adsorbed onto sediment, especially if it is rich in organic matter. Volatilization from soil is also likely, although, depending on the characteristics of the soil, there may also be sorption to soil.

The majority of chlorobenzenes added to soil, as either sewage sludge or spiked samples, were volatilized, with biodegradation and abiotic degradation insignificant compared with the amount volatilized (Wang & Jones, 1994a). Volatilization occurred by two-step first-order processes, with high rates of volatilization during an initial step, followed by a second, much slower step, which was presumably controlled by the rate of desorption of the compound from soil. Half-lives for loss of chlorobenzenes ranged from 13.0 to 219 days for sewage sludge applications and from 10.6 to 103 days for spiked samples. Half-lives increased with increasing chlorination and were also higher in sludge-amended soil than in the spiked samples. The half-lives for volatilization of MCB and 1,2-DCB from soil were 2.1 and 4.0 days, respectively. Initial soil concentrations were 100 mg/kg dry weight (Anderson et al., 1991). Transient geochemical conditions can significantly alter the extent of removal. Robertson (1994) studied the fate of a dichlorobenzene mixture (containing 74% 1,2-DCB, 11% 1,3-DCB, and 15% 1,4-DCB) released to sub-surface soil in effluent from a septic system. High dichlorobenzene concentrations were found in the aerobic unsaturated zone (below the septic system) where dichlorobenzene had a residence time of 60 days. The migration of dichlorobenzene to the water table was attenuated by this zone.

The most important factor affecting the behaviour and fate of chlorobenzenes in soil is sorption. Adsorption–desorption processes in soil affect the rate of volatilization and leaching and the availability of chemicals to microbial and chemical degradation or uptake by plants or other organisms (Wang & Jones, 1994a). The soil sorption coefficients for chlorobenzenes range from 466 to 58 700 (Table 1) and generally increase with increasing chlorination (IPCS, 1991a; Schrap et al., 1994). Sorption of chlorobenzenes to soil is affected by many parameters, and it increases with increasing organic matter content (Barber et al., 1992; Faschan et al., 1993).

The adsorption of 1,2,4-TCB to soil was found to decrease with increasing soil depth (Njoroge et al., 1998). These depth-related changes were attributed to changes in composition, texture, and accessibility of the soil organic matter. At deeper levels, extractable organic matter was increasingly dominated by fulvic acids. The higher fulvic:humic acid ratio in deep soil reflects an inceasing hydrophilicity of the soil organic matter. Abundance of iron oxide and size of clay particles also increase with depth.

Sorption of chlorobenzenes is also affected by soil moisture, with reduced sorption to wet soil (Chiou & Shoup, 1985; Thibaud et al., 1993). Adsorption of 1,2,4-TCB to soil was reduced following the addition of sodium dodecyl sulfate, a surfactant that frequently occurs in sewage sludge disposed of onto land (DiVincenzo & Dentel, 1996). Desorption occurred only when the sodium dodecyl sulfate concentration exceeded the critical micelle concentration. However, increased adsorption of MCB to soil was reported following the addition of the surfactant hexadecyltrimethylamminium (HDTMA) (Sheng et al., 1998). Adsorption on the HDTMA phase was 80–160 times higher than sorption on natural organic matter. The sorption of 1,4-DCB by aquifer materials with a low organic carbon content was enhanced in the presence of tetrachloroethene (Brusseau, 1991). The enhanced sorption was suggested to arise from tetrachloroethene increasing the organic carbon content of the sorbent.

Desorption of 1,3-DCB from a silty soil to deionized water had an initial fast labile phase, followed by a slow phase (Lee et al., 2002). An average of 60% of the initial concentration was desorbed. The first-order rate constant was 0.022–0.038 per hour for the labile phase and 4.1 × 10^–5 to 7.8 × 10^–4 per hour for the slow desorption phase. Single-step batch tests showed that desorption of chlorobenzenes from sediment was slow, with less than 0.5% of 1,2,4,5-TeCB and PeCB desorbed within 62 days. Desorption of 1,2,4-TCB was significantly higher than that of other compounds, with 3% desorbed within 62 days (Gess & Pavlostathis, 1997).

MCB adsorbed onto marine sediment reached equilibrium within 3 h (Zhao et al., 2001). Equilibrium took the same time in natural seawater, artificial seawater, and deionized water. Adsorption occurred via the surface and micropores of sediment and could be described by either the Freundlich or Langmuir model. Adsorption was not affected by temperature (18, 25, or 30 °C), although the saturate adsorption amount decreased at higher temperatures. Adsorption isotherms and the saturate adsorption amounts were higher in natural seawater than in artificial seawater and deionized water. Adsorption of 1,2,4,5-TeCB on sandy aquifer solids took up to hundreds of days to reach equilibrium (Ball & Roberts, 1991). Distribution coefficients were greatest in the size fraction with the largest grains.

Mean (± SD) suspended sediment/water partition coefficients (log K_oc) for chlorobenzenes measured in Ise Bay, Japan, were 3.47 ± 0.74 (1,3-DCB), 3.69 ± 0.48 (1,2-DCB), 3.61 ± 0.39 (1,2,3-TCB), 3.86 ± 0.40 (1,2,4-TCB), 3.55 ± 0.47 (1,3,5-TCB), 4.39 ± 0.33 (1,2,3,4-TeCB), 3.94 ± 0.33 (1,2,3,5-TeCB and 1,2,4,5-TeCB), and 4.59 ± 0.41 (PeCB) (Masunaga et al., 1996b). Concentrations of chlorobenzenes in water and adsorbed onto suspended sediment were compared. None of the chlorobenzenes gave a clear adsorbed level distribution pattern, and the correlation between soluble and adsorbed chlorobenzenes was weak.

The fate of MCB, 1,2-DCB, and 1,2,4-TCB in wastewater applied to soil was examined in a microcosm experiment (Piwoni et al., 1986). Initial concentrations of MCB, 1,2-DCB, and 1,2,4-TCB in the wastewater were 1.9–3.1, 2.4, and 0.72 µmol/litre, respectively. The proportions of MCB and 1,2-DCB volatilized were 14% and 21%, respectively, and it was assumed that 84% and 79%, respectively, were degraded, giving concentrations in the volume effluent of 9 ± 10% of the original concentration. Volatility of 1,2,4-TCB was not measured, but it was assumed to be approximately 89%, as <0.7% of the original concentration remained in the effluent.

The half-life for dichlorobenzene (all isomers) in a septic groundwater system was 15 days (Robertson, 1994). The site included a 2-m-thick, sandy aerobic unsaturated zone. This loss was due to a combination of volatilization and aerobic biodegradation. Biodegradation occurred after an initial lag phase and was most likely for 1,3-DCB and 1,4-DCB. Dichlorobenzene in the anaerobic zone was not readily biodegraded.

Octanol/air partition coefficients (log K_oa) measured for chlorobenzenes at 25 °C were 4.36 (1,2-DCB), 5.19 (1,2,3-TCB), 5.64 (1,2,3,4-TeCB), 5.63 (1,2,4,5-TeCB), and 6.27 (PeCB) (Harner & Mackay, 1995). Octanol/air partition coefficients determined partitioning from the atmosphere to vegetation, soils, and possibly aerosols.

Microcosm experiments suggested that 1,2-DCB in soil was not taken up by grass (Holcus lanatus) roots, although some foliar adsorption of dichlorobenzene volatilized from soil was reported (Wilson & Meharg, 1999). A root concentration factor of 19 litres/kg has been reported for 1,2,4-TCB (Dietz & Schnoor, 2001). From these data, it can be assumed that tri- and/or tetrachlorinated benzenes have the potential to be taken up by plants.

5.2.1 Abiotic degradation

Removal of chlorobenzenes from the atmosphere will occur primarily via reactions with hydroxyl radicals to produce nitrochlorobenzene, chlorophenol, and aliphatic dicarbonyl products, which are further removed by photolysis or reaction with hydroxyl radicals. Photolysis and reactions with ozone or nitrate radicals are of negligible importance (Grosjean, 1991). Rate constants for reactions with hydroxyl radicals (in cm³/s per molecule) were calculated to be 8.8 × 10^–13 (MCB), 4.0 × 10^–13 (1,2-DCB), 7.2 × 10^–13 (1,3-DCB), 4.3 × 10^–13 (1,4-DCB), 6.0 × 10^–13 (1,2,3-TCB), and 5.65 × 10^–13 (1,2,4-TCB) (Atkinson et al., 1985; Klöpffer et al., 1986; Dilling et al., 1988; Arnts et al., 1989). A rate constant for reaction of MCB with ozone was calculated to be <5 × 10^–21 cm³/s per molecule. Assuming 24-h average hydroxyl radical and ozone concentrations of 1 × 10⁶ and 7.2 × 10¹¹ molecules/cm³, tropospheric half-lives for MCB were calculated to be 13 days for reactions with hydroxyl radicals and >8.8 years for reactions with ozone (Atkinson et al., 1985). Tropospheric half-lives for 1,4-DCB and 1,2,4-TCB reacting with hydroxyl radicals were calculated to be 33.4 and 26.7 days, respectively (Klöpffer et al., 1988).

1,2,4-TCB in the atmosphere may be degraded via direct photolysis, although this route of degradation is minor, due to the poor spectral overlap between the solar spectrum and the adsorption spectrum of 1,2,4-TCB. The maximum photolysis rate for 1,2,4-TCB in summer at midday under clear skies was 0.03% per hour (Bunce et al., 1989).

Chlorobenzenes in aqueous solutions may undergo photochemical reductive dechlorination. PeCB was degraded to tetrachlorobenzenes, which in turn were photodegraded to trichlorobenzenes, dichlorobenzenes, MCB, and, ultimately, phenol, benzene, and hydrogen chloride (Chu & Jafvert, 1994). These reactions were reported following exposure to 253.7-nm monochromatic ultraviolet lamps. The rate of photodegradation increased in the presence of surfactants. In addition to the main reductive pathway of photodechlorination, minor pathways, including photochlorination, photohydrolysis, and photoisomerization, also occurred. 1,2,3,5-TeCB was photolysed to 1,2,4-TCB or 1,3,5-TCB in the presence of an acetone sensitizer (Choudhry & Hutzinger, 1984). Photochemical reactions in the absence of a sensitizer transformed tetrachlorobenzenes into other isomers and also produced some chlorobenzenes with greater chlorination than the original tetrachlorobenzene compound. The rate constant for reaction of 1,2,4-TCB with hydroxyl radicals in an acidic solution was 6.0 ± 0.3 × 10⁹ per mol/litre per second (Gallard & De Laat, 2001). 1,4-DCB in aqueous solution was photodegraded to 4-chlorophenol, hydroquinone, hydroxybenzoquinone, and 2,5-dichlorophenol (Meunier et al., 2001). The formation of 2,5-dichlorophenol demonstrates hydroxylation without dechlorination. Photodegradation of MCB in aqueous solutions has been reported under both aerobic and anaerobic conditions and at pHs ranging from <1 to <12 (Tissot et al., 1983, 1984; Dilmeghani & Zahir, 2001). Degradation followed first-order kinetics, with rate constants ranging from 1.8 × 10^–4 to 6.4 × 10^–4 per second for anaerobic and oxygen-saturated conditions, respectively (Dilmeghani & Zahir, 2001). The rate of degradation was an order of magnitude higher with ultraviolet and hydrogen peroxide or hydrogen peroxide–ozone compared with ultraviolet alone.

The half-lives for photolytic degradation of MCB and 1,2,4-TCB in surface water, simulating summer conditions at 40 degrees latitude, were 170 and 450 years, respectively (Dulin et al., 1986).

5.2.2 Biodegradation

Chlorobenzenes in various substrates, including soil, sediment, and sewage sludge, can be degraded by microorganisms. The major mechanism of aerobic degradation is via oxidative dechlorination, usually initiated by dioxygenative hydroxylation, leading to the formation of hydroxylated aromatic compounds (mainly catechols), which undergo ring fission and subsequent mineralization to carbon dioxide and water. The less chlorinated benzenes are more readily degraded than the higher chlorinated ones (IPCS, 1991a). Biodegradation under anerobic conditions has also been reported, although this occurs at a slower rate than aerobic biodegradation.

Chlorobenzene-degrading bacteria isolated from aerobic environments include Burkholderia (previously known as Pseudomonas) species (strains JS150, P51, JS6, PS12, and PS14) (Pettigrew et al., 1991; Sander et al., 1991; Van der Meer et al., 1991, 1997; Nishino et al., 1994; Beil et al., 1997; Meckenstock et al., 1998), Alcaligenes species (strains A175 and OBB65) (De Bont et al., 1986; Schraa et al., 1986), Escherichia hermanii (Kiernicka et al., 1999), Nitrosomonas europaea (Keener & Arp, 1994), Mycobacterium vaccae, and Rhodococcus species (strain R22) (Fairlee et al., 1997).

The degradative abilities of these bacteria vary, with some organisms exhibiting a lag or adaptation period prior to degradation. Some can degrade several chlorobenzenes (Brunsbach & Reineke, 1994), whereas others are compound-specific (Reineke & Knackmuss, 1984; Brunsbach & Reineke, 1994; Keener & Arp, 1994). For some, degradation occurs only in the presence of other sources of carbon and energy, whereas others are able to use chlorobenzenes as their sole carbon and energy source (Van der Meer et al., 1987). Genetic analysis has shown that these bacteria contain a novel combination of previously existing genes — genes for aromatic ring dioxygenase and dihydrodiol dehydrogenase — and other genes for a chlorocatechol oxidative pathway.

Degradation is also dependent upon the initial chlorinated benzene concentrations. Degradation will occur only if the initial concentration is below the toxic threshold, although bacteria that have previously been exposed to MCB have the ability to degrade higher concentrations than those that did not have prior exposure. For example, concentrations of MCB greater than 2.5 mmol/litre (282 mg/litre) were found to be toxic to Pseudomonas sp. strain RHO1 cells. Cells that had previously been exposed to MCB demonstrated toxicity at concentrations greater than 3.5 mmol/litre (394 mg/litre) (Fritz et al., 1992).

MCB and 1,2,4-TCB were degraded by bacteria isolated from solids sampled from pristine aquifers (Swindoll et al., 1988). Degradation followed first-order rate constants, with V_max values of 0.38–2.71 ng/g per hour for 1,2,4-TCB. Degradation of MCB was not saturated; therefore, V_max could not be calculated.

A consortium of Gram-negative and Gram-positive bacteria isolated from groundwater and soil contaminated with MCB was able to mineralize 54% of a 2.23 µmol/litre solution via the modified ortho pathway within 7 days in the presence of nutrients. Degradation also occurred without added nutrients, although at a slower rate (Nishino et al., 1992).

Degradation of 1,4-DCB occurred at similar rates under aerobic or anaerobic conditions and was enhanced in mixtures with high sludge content (which reduced overall oxygen) (Gejlsbjerg et al., 2001). Mineralization occurred after a lag phase of 30 days. After inoculation for 2 months, mineralization was 12.4% in the sludge and 21.6% in the 1:20 sludge:soil mixture. The authors concluded that mineralization was probably occurring in the aerobic layers of the sludge–soil mixtures, as mineralization did not occur in sludge in the absence of molecular oxygen.

A consortium of bacteria isolated from Rhine sediment was able to degrade PeCB, 1,2,3,4-TeCB, 1,2,3,5-TeCB, 1,2,4,5-TeCB, and 1,2,3-TCB via reductive dechlorination in the presence of lactate, glucose, ethanol, or isopropanol as the electron donor (Holliger et al., 1992). PeCB was degraded to 1,3,5-TCB, while 1,2,3,4-TeCB and 1,2,4,5-TeCB were degraded to 1,2,4-TCB. Chlorobenzenes that were not dechlorinated during the 4-week incubation included 1,2,4-TCB, 1,3,5-TCB, and all isomers of dichlorobenzene. Other studies have reported complete mineralization of some higher chlorinated compounds. Two Pseudomonas strains (PS12 and PS14) isolated from the soil of an industrial waste deposit were able to mineralize various chlorobenzenes, including MCB, all three dichlorobenzenes, 1,2,4-TCB, and 1,2,4,5-TeCB (strain PS14 only). 1,2,4-TCB and 1,2,4,5-TeCB were degraded via dioxygenation of the aromatic ring, producing 3,4,6-trichlorocatechol. Subsequent ortho cleavage, catalysed by a Type II catechol 1,2-dioxygenase, produced 2,3,5-trichloromuconate, which was degraded via the tricarboxylic acid pathway (Sander et al., 1991).

Degradation of 1,2-DCB and 1,4-DCB within a mixture of organic compounds was reported in a 149-day batch microcosm using sediment and groundwater obtained from various sampling sites of an aquifer (Nielsen & Christensen, 1994). The initial concentrations were approximately 120 µg/litre. Within an average of 82 days, 78.3% of 1,4-DCB and 81.0% of 1,2-DCB were degraded. The lag phases were 4.9 and 4.5 days for 1,4-DCB and 1,2-DCB, respectively. In the Organisation for Economic Co-operation and Development (OECD) closed bottle test, 67% of an initial 1,4-DCB concentration of 1.9 mg/litre was mineralized after 28 days, indicating that 1,4-DCB is readily degradable (Topping, 1987).

Bartholomew & Pfaender (1983) calculated degradation rates for MCB and 1,2,4-TCB at different sites of a river system during different seasons. Rates of degradation of MCB and 1,2,4-TCB were reported to decrease over the freshwater to estuarine to marine gradient. V_max values for MCB degradation during May and September were 13–14 ng/litre per hour for fresh water, 4.9–10 ng/litre per hour for estuarine water, and <1–1.7 ng/litre per hour for marine water. V_max values were <1 ng/litre per hour at all three sites in February. The corresponding values for degradation of 1,2,4-TCB in May and July were <1–7.5 ng/litre per hour for fresh water, <1–7.9 ng/litre per hour for estuarine water, and <1–2.3 ng/litre per hour for marine water.

In controlled lysimeter experiments, 80% of 1,2,4,5-TeCB in soils and liquid cultures was mineralized by the bacterial strains Isphingomonas sp. strains HH69 and RW1 and Pseudomonas sp. strain PS14 within a few days (Figge et al., 1993). Degradation was not increased in the presence of additional energy sources such as peptone, triolein, and glucose. Degradation did not occur in acidic soils (pH < 4).

Biodegradation of chlorobenzenes has also been reported in several studies under anaerobic conditions, including methanogenic and sulfate-reducing conditions. As with aerobic degradation, degradability varies between organisms. Under anaerobic conditions, degradation is limited to dechlorination, with no breakdown of the aromatic structure.

Anaerobic degradation of chlorobenzenes has been reported in river sediment (Masunaga et al., 1996a; Susarla et al., 1996). Dechlorination occurred without a lag period, with half-lives ranging from 17 to 433 days. The main pathway for PeCB dechlorination was via 1,2,4,5-TeCB, 1,2,4-TCB, 1,4-DCB, and MCB. A minor pathway, via 1,2,3,4-TeCB, 1,2,3-TCB, 1,2-DCB, and MCB, was also observed. MCB was stable under anaerobic conditions. The preferences for dechlorination were two adjacent chlorine atoms, followed by one chlorine on an adjacent carbon, followed by no chlorine on the adjacent carbon. Other studies have reported similar anaerobic biodegradation (Beurskens et al., 1991; Ramanand et al., 1993; Susarla et al., 1997). Nowak et al. (1996) reported anaerobic degradation of all chlorobenzenes, including MCB, to benzene.

In anaerobic sewage sludge, PeCB was dechlorinated to 1,2,3,4-TeCB and 1,2,3,5-TeCB, which were degraded to 1,2,4-TCB, 1,2,3-TCB, and 1,3,5-TCB, and then 1,2-DCB and 1,3-DCB (Yuan et al., 1999). Sequential dechlorination occurred within a substrate concentra-tion range of 2–50 mg/litre, but was slower at concentrations greater than 50 mg/litre. Dechlorination rates were highest under methanogenic conditions (0.30 mg/litre per day), with slower rates under sulfate-reducing (0.12 mg/litre per day) and denitrifying conditions (0.08 mg/litre per day). The rate of dechlorination of 1,2,3-TCB by anaerobic sediment ranged from 15 to 35 pmol/ml wet sediment per day (Yonezawa et al., 1994).

Some studies have shown chlorobenzenes to be resistant to anaerobic biodegradation. Nielsen et al. (1995) reported no biodegradation of 1,2-DCB or 1,4-DCB in anaerobic landfill leachate collected from four different sites at distances ranging from 2 to 350 m from the landfill. The governing reactions, which varied at each site, included methanogenesis, iron(III) reduction, nitrate reduction, and manganese(IV) reduction. Dichlorobenzenes have been reported to persist for at least 20 years in an aquifer that had been contaminated with rapid-infiltration sewage disposal (Barber, 1988). 1,2,3,5-TeCB and 1,3,5-TCB were resistant to degradation by soil slurry microorganisms that could degrade PeCB, 1,2,3,4-TeCB, and 1,2,4-TCB (Ramanand et al., 1993).

The bioaccumulation of chlorobenzenes by aquatic organisms is determined by their relative water and lipid solubility (thus reflecting the octanol/water partition coefficients) and the number of chlorine substitutions. Uptake from water increases with increasing chlorination (Könemann & Van Leeuwen, 1980; Oliver & Niimi, 1983; Sabljic, 1987; Koelmans & Jimenez, 1994; Wang et al., 1997) and with increasing temperature (Koelmans & Jimenez, 1994).

Mean bioconcentration factors (BCFs) (dry weight) for phytoplankton increased from 4700 for 1,2,3-TCB at 4.5 °C to 26 000 for PeCB at 38.6 °C (Koelmans & Jimenez, 1994). Wang et al. (1997) found significant differences in the accumulation of chlorobenzenes by different marine algal species, with BCFs (dry weight) ranging from 600 to 3000 for 1,2,3,4-TeCB and from 1000 to 6000 for PeCB.

BCFs ranging from 270 for 1,2-DCB to 20 000 for PeCB were reported for laboratory studies on rainbow trout (Oncorhynchus mykiss) (Oliver & Niimi, 1983). BCFs for a variety of fish species ranged from 7000 to 24 000 (lipid weight) for 1,2,4-TCB, with a positive correlation between bioaccumulation and lipid content (Geyer et al., 1985). Galassi & Calamari (1983) found BCFs (lipid weight) ranging from 4000 to 22 000 for 1,2,3- and 1,2,4-TCB in rainbow trout, with newly hatched fish accumulating 2–4 times the amount found in eyed eggs or young fish (alevins). Qiao et al. (2000) report that gill uptake of 1,2,4-TCB and PeCB could account for 98% of the body burden. Uptake of trichlorobenzenes, tetrachlorobenzenes, and PeCB was significantly reduced by the presence of suspended particles (Schrap & Opperhuizen, 1990). However, PeCB was found to be readily desorbed from sediments with a low organic carbon content and subsequently accumulated by fish via the gills (Qiao & Farrell, 1996). The rate of elimination of chlorobenzenes decreases with increasing chlorination (Melancon & Lech, 1985; De Boer et al., 1994). Elimination half-lives for dichlorobenzenes to PeCB in laboratory-exposed fish ranged from 0.05 to 1.6 days (Melancon & Lech, 1985). However, for eels (Anguilla anguilla) transferred from a contaminated lake to a "clean" lake, elimination half-lives of >300 days were reported for tetrachlorobenzenes and PeCB (De Boer et al., 1994). Sijm & Van der Linde (1995) calculated elimination rate constants and predicted elimination half-lives for 1,2,3-TCB to be 40 days in small fish, such as guppies (Poecilia reticulata), and >5 years in larger and/or fatty fish.

The coefficient of adsorption onto sediment influences the uptake into terrestrial plants and sediment-living aquatic invertebrates; the degree of chlorination is also correlated with uptake (Knezovich & Harrison, 1988; IPCS, 1991a). Under non-equilibrium conditions, BCFs for chironomid midge larvae exposed to sediment-bound chlorobenzenes were 5, 29, and 225 for MCB, 1,2-DCB, and 1,2,4-TCB, respectively. BCFs were best correlated with the concentrations of the chlorobenzenes in the interstitial water (Knezovich & Harrison, 1988).

The tri- and tetrachlorinated benzenes may be taken up by plants, as indicated by the root concentration factor of 19 litres/kg reported for 1,2,4-TCB (Dietz & Schnoor, 2001).

However, the prediction of BCFs is more difficult for terrestrial plants than for aquatic organisms because of the complex nature of the root soil interface combined with gaseous uptake by aerial parts (Scheunert et al., 1994). Topp et al. (1986) compared the uptake of chlorobenzenes by plants from the soil and via the air in closed, aerated laboratory systems. A negative correlation was demonstrated between the BCF and the soil adsorption coefficient (based on soil organic matter content) for the uptake into the roots of barley. The adsorption of chlorobenzenes onto soil organic matter increased with increasing chlorination. However, expression of uptake in barley roots in relation to the soil interstitial water concentration of the chlorobenzenes produced a positive correlation between the BCF and the octanol/water partition coefficients. Higher chlorinated chlorobenzenes, therefore, are most readily taken up by the plant roots when they are available in soil interstitial water. This will occur particularly in sandy soils with low organic matter content. In a later study, Topp et al. (1989) found that after growth in soil containing 2 µg each of 1,2,4-TCB and PeCB per kg dry weight, harvested barley grain contained 73 and 82 µg/plant, respectively. The concentrations in the dry grain were 0.05 and 0.06 mg/kg for 1,2,4-TCB and PeCB, respectively. In further studies on soybeans (Glycine max), linear correlations were found between equilibrium tissue/water coefficients, the octanol/water partition coefficient, and measured lipid content (Tam et al., 1996). The bioconcentration of chlorobenzenes into excised soybean (Glycine max) roots increased exponentially with increasing octanol/water partition coefficient (Kraaij & Connell, 1997). Wang & Jones (1994b) concluded that the total amount of chlorobenzenes taken up by carrots grown in sewage sludge-amended and spiked soils was low (<1%) compared with other loss pathways from the soil, principally volatilization.

Belfroid et al. (1994) calculated BCFs for earthworms (Eisenia andrei) of 104 and 156 for 1,2,3,4-TeCB and PeCB in soil; BCFs based on interstitial water were 67 000 and 307 000, respectively, and were found to be similar to BCFs found for worms exposed in water alone (Belfroid et al., 1993). BCFs for earthworms exposed via water show a clear increase in uptake of chlorobenzenes with increasing chlorination, and steady-state concentrations are reached within 5 days (Belfroid et al., 1993). Elimination rate constants reveal that chlorobenzene loss decreases with increasing chlorination. A monophasic elimination curve was observed in water, whereas biphasic elimination was found in the presence of soil (Belfroid et al., 1993); elimination rates in soil experiments were significantly increased by the addition of organic matter (Belfroid & Sijm, 1998). Feeding studies have revealed that earthworms can also take up chlorobenzenes via food. In studies with field-contaminated soil, steady-state concentrations in worms were much lower than in laboratory studies, suggesting decreased bioavailability of chlorobenzenes (Belfroid et al., 1995).

Chlorobenzene (MCB, dichlorobenzenes, and trichlorobenzenes) concentrations have previously been reported in ambient air, with mean concentrations in the order of 0.1 µg/m³ and maximum levels of up to 100 µg/m³ at hazardous waste sites (IPCS, 1991a). Popp et al. (2000) measured tetrachlorobenzenes and PeCB in air sampled from two industrially contaminated sites and a reference site in Germany in 1998. Mean gas-phase concentrations of tetrachlorobenzenes and PeCB at the contaminated sites ranged from 5.7 to 30.9 pg/m³ and from 10.2 to 28 pg/m³, respectively. Mean concentrations at the control site ranged from 6.4 to 10.6 pg/m³. Particulate-bound chlorobenzenes accounted for 1.9% of the total concentrations. A low proportion of particulate-bound chlorobenzenes was also reported in air sampled from the Bering and Chukchi seas in 1993 (Strachan et al., 2001). Mean gas-phase concentrations for the Bering Sea were 1.1, 4.0, and 6.6 pg/m³ for 1,2,3-TCB, 1,2,3,4-TeCB, and PeCB, respectively, and for the Chukchi Sea, 2.8, 10, and 14 pg/m³, respectively. Mean chlorobenzene concentrations at four sites throughout Michigan, USA (1992–1994), ranged from 22 to 30 pg/m³ for 1,2,4,5-TeCB, from 40 to 53 pg/m³ for 1,2,3,4-TeCB, and from 35 to 69 pg/m³ for PeCB (Hermanson et al., 1997). Annual mean concentrations for southern Ontario, Canada (1988–1989), were >5.3 pg/m³ for 1,2,3,4-TeCB and >8.0 pg/m³ for PeCB (Hoff et al., 1992). Higher concentrations have been reported in close proximity to pollution sources. A concentration of 5 µg/m³ for tri- and tetrachlorobenzenes was found within 200 m of an electro-industrial plant in Slovenia (Jan et al., 1994). Seasonal variations in the concentrations of 1,4-DCB in ambient air have also been reported, with concentrations increasing with increasing temperature (Hanai et al., 1985).

Chlorobenzenes have also been detected in rainwater, their presence presumably being due to transfer from the ambient air. Concentrations of all three dichlorobenzene isomers and 1,2,4-TCB in rainwater were less than 10 ng/litre at selected sites in Oregon and California, USA (Pankow et al., 1983). In the United Kingdom, 1,4-DCB was detected in rainwater at a mean concentration of 10 ± 5 ng/litre (Fielding et al., 1981).

In 12 sewage sludges in the United Kingdom, the concentrations of chlorobenzenes ranged from <0.01 mg/kg dry weight for PeCB to 40.2 mg/kg dry weight for 1,3-DCB, with a general reduction in concentration with increased chlorine substitution (Rogers et al., 1989). Further sampling of United Kingdom sewage sludges revealed chlorobenzene concentrations ranging from 35 100 to 192 000 mg/kg dry weight for MCB, from 13 to 4110 mg/kg for dichlorobenzenes, from 2 to 1070 mg/kg for trichlorobenzenes, from 0.2 to 101 mg/kg for tetrachlorobenzenes, and from 2 to 37 mg/kg for PeCB (Wang & Jones, 1994c). Analysis of archived sludge samples showed that concentrations of 1,4-DCB increased over the period 1942–1961, whereas other chlorobenzenes increased in concentration only from 1954 onwards (Wang et al., 1992).

Data on levels of the lower chlorinated benzenes (MCB, dichlorobenzenes, and trichlorobenzenes) in wastewater indicate that MCB is detected the most often and at the highest concentrations, occasionally exceeding 1 mg/litre. Chlorobenzene concentrations in US wastewater have been reported to range from 11 to 6400 µg/litre for MCB, from 10 to 860 µg/litre for dichlorobenzenes, and from 12 to 607 µg/litre for trichlorobenzenes (IPCS, 1991a).

Concentrations of chlorobenzenes in surface waters are generally in the ng/litre to µg/litre range, with maximum concentrations up to 0.2 mg/litre in areas close to industrial sources (IPCS, 1991a). Mean concentrations of dissolved chlorobenzenes in the Bering and Chukchi seas ranged from 3 to 10 pg/litre for 1,2,3-TCB, from 15 to 36 pg/litre for 1,2,3,4-TeCB, and from 9 to 36 pg/litre for PeCB (Strachan et al., 2001). Higher chlorobenzene levels have been detected in coastal waters and estuaries, with Dutch coastal waters containing mean concentrations ranging from 9 to 117 ng/litre for dichlorobenzenes and from 0.7 to 1.6 ng/litre for trichlorobenzenes (Van de Meent et al., 1986) and Japanese coastal waters containing mean dissolved concentrations ranging from 24.3 ng/litre for 1,3-DCB to 0.25 ng/litre for tetrachlorobenzenes (Masunaga et al., 1996b). Waters of the Scheldt estuary (The Netherlands) contained chlorobenzene concentrations ranging from <130 to 315 ng/litre for dichlorobenzenes, from <25 to 320 ng/litre for trichlorobenzenes, and from <45 to 135 ng/litre for tetrachlorobenzenes (Van Zoest & Van Eck, 1991); more recent sampling revealed MCB concentrations ranging from 5 to 31.5 ng/litre (Huybrechts et al., 2000). Chlorobenzene concentrations of up to 500 ng/litre have been reported for MCB in the Tees Estuary, United Kingdom (Law et al., 1991), and for 1,3-DCB in Yokkaichi Port, Ise Bay, Japan, during 1988 (Masunaga et al., 1991a). Mean chlorobenzene concentrations in the Forth Estuary, United Kingdom, during 1987 ranged from <0.1 to 790 ng/litre for dichlorobenzenes, from 4 to 5500 ng/litre for trichlorobenzenes, from <0.04 to 20 ng/litre for tetrachlorobenzenes, and from <0.01 to 40 ng/litre for PeCB. The predominant isomers detected were 1,2,3- and 1,2,4-TCB, and these were found near industrial effluent discharges (Rogers et al., 1989; Harper et al., 1992). Further studies in 1990 revealed 1,2,3- and 1,2,4-TCB concentrations ranging up to 51 and 84 ng/litre, respectively (Harper et al., 1992).

The highest chlorobenzene concentrations in surface waters have been reported for river waters in heavily populated and/or industrialized areas. Mean concentrations in the river Besos, Spain, were 260 ng/litre for MCB, 600 ng/litre for 1,4-DCB, 5000 ng/litre for 1,2-DCB and 1,3-DCB, 1100 ng/litre for 1,2,3-TCB, and 8100 ng/litre for 1,2,4-TCB (Gomez Belinchon et al., 1991). Concentrations of MCB and 1,4-DCB ranging from non-detected to >10 µg/litre have been reported for both compounds in water from the Ohio River (US EPA, 1985). Elder et al. (1981) reported trichlorobenzene concentrations (isomer not specified) ranging from 0.1 to 8 µg/litre in water from Niagara Falls, New York, USA. Corresponding concentrations of tetrachlorobenzene ranged from 0.1 to 200 µg/litre. Concentrations of PeCB in water sampled from the Great Lakes ranged from not detected to 0.0006 µg/litre (Oliver & Nicol, 1982). Concentrations in water sampled from the rivers and estuary of Osaka (a major urban area of Japan) ranged from 0.2 to 30 µg/litre for MCB, from 0.17 to 130 µg/litre for 1,4-DCB, from 0.2 to 10 µg/litre for 1,2-DCB, from 0.16 to 0.35 µg/litre for 1,2,4-TCB, and from 0.18 to 0.30 µg/litre for 1,2,3-TCB (Yamamoto et al., 1997).

Mean chlorobenzene concentrations in sediment from the Bering and Chukchi seas ranged from 0.02 to 0.41 µg/kg for 1,2,3-TCB, from 0.08 to 0.87 µg/kg for 1,2,3,4-TeCB, and from 0.33 to 0.4 µg/kg for PeCB (Strachan et al., 2001). Mean concentrations in coastal sediments from Ise Bay, Japan, were 4.8 µg/kg for 1,2,4-TCB, 2.3 µg/kg for 1,2-DCB, 1.9 µg/kg for 1,3-DCB, and <0.15 µg/kg for 1,3,5-TCB, tetrachlorobenzenes, and PeCB (Masunaga et al., 1991b). Lee & Fang (1997) reported mean values for the Tsen-wen estuary, Taiwan, of 3.2 µg/kg for 1,2-DCB, 20.7 µg/kg for 1,3-DCB, and 11.2 µg/kg for 1,2,4-TCB.

Lake Garda (Italy) contained mean sediment PeCB concentrations of 0.2 µg/kg dry weight (Bossi et al., 1992), whereas Lake Superior (Hamilton Harbour, Canada) contained levels ranging from 3.6 µg/kg for PeCB to 80 µg/kg for 1,4-DCB (Onuska & Terry, 1985). Sediment samples from the river Elbe, Germany, ranged from 30 to 740 µg/kg dry weight for MCB, from 20 to 1060 µg/kg for dichlorobenzenes (1,2- and 1,4-DCB), from 1 to 115 µg/kg for trichlorobenzenes (1,2,3- and 1,2,4-TCB), from 1 to 27 µg/kg for tetrachlorobenzenes, and from 1 to 14 µg/kg for PeCB (Götz et al., 1993), whereas samples from the river Rhine contained concentrations ranging from 40 to 240 µg/kg dry weight for dichlorobenzenes, from <10 to 20 µg/kg for trichlorobenzenes, and from <0.5 to 2 µg/kg for PeCB (Alberti, 1983).

Chlorobenzene levels in uncontaminated soils are generally less than 0.4 mg/kg for dichlorobenzene congeners and less than 0.1 mg/kg for other chlorobenzene congeners (Wang et al., 1995). Multiple applications of sewage sludge can increase the chlorobenzene content in sludge-amended soil compared with control soils. However, Wang et al. (1995) found that most chlorobenzenes disappear rapidly on cessation of sludge application, with around 10% remaining 30 years later. They found that 1,4-DCB levels increased significantly in United Kingdom soils during the 1960s to a maximum mean value in 1967 of 10 mg/kg in control soils and 16.6 mg/kg in sludge-amended soils. Analysis of subsoil from a former pesticide factory in Germany showed that tetrachlorobenzenes and PeCB were dominant in the upper soil layers (up to 1.9 m), accounting for 80% of chlorobenzenes, with 1,2,3,4-TeCB and PeCB accounting for 44% and 24%, respectively. At depths between 1.9 and 5.5 m, trichlorobenzenes were more dominant, accounting for 60%, with 1,2,4-TCB accounting for 37% (Feidieker et al., 1994). Total chlorobenzene concentrations ranged from 1.5 to 18 400 mg/kg.

Mean chlorobenzene concentrations in bivalves from US coastal waters ranged from <0.25 to 28.2 µg/kg dry weight for 1,2,4,5-TeCB, from <0.25 to 10 µg/kg for 1,2,3,4-TeCB, and from <0.25 to 13.3 µg/kg for PeCB (Wade et al., 1998). Aquatic insects from a variety of Canadian sites contained mean PeCB concentrations ranging from <0.49 to 21.4 µg/kg dry weight (Ciborowski & Corkum, 1988). Concentrations in freshwater and marine fish from contaminated areas range from 0.1 to 50 µg/kg wet weight, with higher chlorinated compounds generally present at the highest concentrations (IPCS, 1991a). The eggs of fish-eating birds contained mean PeCB levels of 1.2 and 4.4 µg/kg from two sites in Puget Sound, USA (Cobb et al., 1994). Waterfowl from Lake Ontario, Canada, contained mean chlorobenzene concentrations ranging from 0.3 to 1.7 µg/kg wet weight for 1,2,3,4-TeCB and from 0.65 to 33.4 µg/kg for 1,2,4,5-TeCB (Gebauer & Weseloh, 1993). Mean concentrations in Arctic marine mammal blubber ranged from 1 to 9.7 µg/kg wet weight for 1,2,3,4-TeCB and from 16.8 to 20.2 µg/kg for PeCB (Muir et al., 1992; Weis & Muir, 1997).

The acute toxicity of chlorobenzenes to aquatic organisms is presented in Table 3. Forty-eight-hour EC₅₀s for diatoms range from 8 to 235 000 µg/litre. For freshwater invertebrates, 48-h EC₅₀s range from 10 µg/litre for PeCB to >530 000 µg/litre for 1,2,4,5-TeCB. Ninety-six-hour LC₅₀s for fish range from 135 for PeCB in the freshwater guppy (Poecilia reticulata) to 21 000 µg/litre for 1,2,4-TCB in the saltwater sheepshead minnow (Cyprinodon variegatus).

Table 3: Acute toxicity of chlorobenzenes to aquatic species.