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 55

POLYCHLORINATED BIPHENYLS:
HUMAN HEALTH ASPECTS

First draft prepared by Dr Obaid M. Faroon, Mr L. Samuel Keith, Ms Cassandra Smith-Simon, and Dr Christopher T. De Rosa, Agency for Toxic Substances and Disease Registry, Atlanta, Georgia, USA

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, 2003

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

Polychlorinated biphenyls : Human health aspects.

(Concise international chemical assessment document ; 55)

1.Polychlorinated biphenyls - toxicity 2.Polychlorinated biphenyls - adverse effects

3.Risk assessment 4.Environmental exposure I.International Programme on Chemical

Safety II.Series

ISBN 92 4 153055 3         (LC/NLM Classification: QV 633)

ISSN 1020-6167

©World Health Organization 2003

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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 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., EHC or 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 based on an existing national, regional, or international review. 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.

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.

The Agency for Toxic Substances and Disease Registry, Division of Toxicology, prepared this CICAD on polychlorinated biphenyls (PCBs) based on the updated Toxicological profile for polychlorinated biphenyls (PCBs) (ATSDR, 2000). In addition, several articles based on the source document can be consulted for details on each of several health end-points considered important in this CICAD (Faroon et al., 2000, 2001a,b). Information on the nature of the peer review and the availability 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 Ottawa, Canada, from 29 October to 1 November 2001. Participants at the Final Review Board meeting are listed in Appendix 3. The International Chemical Safety Card (ICSC 0939) for polychlorinated biphenyl (Aroclor 1254), produced by the International Programme on Chemical Safety (IPCS, 2000), has also been reproduced in this document.

PCBs are synthetic chlorinated hydrocarbon compounds that consist of two benzene rings linked by a single carbon–carbon bond, with from 1 to all 10 of the hydrogen atoms replaced with chlorines. PCBs have been produced commercially since 1929. They have been used in plasticizers, surface coatings, inks, adhesives, flame retardants, pesticide extenders, paints, and microencapsulation of dyes for carbonless duplicating paper. Because PCBs resist both acids and alkalis and are relatively heat-stable, they have been used in dielectric fluids in transformers and capacitors. Further environmental contamination may occur from the disposal of old electrical equipment containing PCBs. The pyrolysis of PCB mixtures produces hydrogen chloride and polychlorinated dibenzofurans (PCDFs), and pyrolysis of mixtures containing chlorobenzenes also produces polychlorinated dibenzodioxins (PCDDs). Many countries have severely restricted or banned the production of PCBs.

The more highly chlorinated PCB congeners adsorb strongly to soil and sediment and are generally persistent in the environment. The various congeners in soil and sediment have half-lives that extend from months to years. Adsorption of PCBs generally increases with the extent of chlorination of the congener and with the organic carbon and clay contents of the soil or sediment. Volatilization and biodegradation — two very slow processes — are the major pathways of PCB removal from water and soil.

PCBs accumulate in the food-chain. They are rapidly absorbed from the gastrointestinal tract and distribute to and accumulate in the liver and adipose tissue. They also cross the placenta, are excreted in milk, and accumulate in the fetus/infant. PCBs are metabolized by hydroxylation and subsequent conjugation. The rates of metabolism and subsequent excretion vary markedly between different congeners.

For the purpose of this CICAD, the health end-points and risk characterization associated with PCB exposures have been based on the approach for mixtures. This is justified on the basis that populations in the general and occupational environments are commonly exposed to mixtures of PCBs, the components of which have different modes of action. It is recognized that in some cases, the mixtures to which various populations are exposed differ considerably from those on which this assessment is based. In such cases, it may be more appropriate to adopt a toxic equivalence (TEQ) approach for individual congeners for which modes of action are known to be similar. Another alternative approach is the use of total body burden of PCB mixtures, since it is done on humans rather than laboratory animals or in vitro, so as to eliminate the need for species extrapolation. Additional information on the various approaches can be found in the source document.²

Humans may be exposed to PCBs by inhaling contaminated air and ingesting contaminated water and food. In 1978, the estimated dietary intake of PCBs by adults in the USA was 0.027 µg/kg body weight per day, but it declined to 0.0005 µg/kg body weight per day in 1982–1984 and <0.001 µg/kg body weight per day for the period 1986–1991.

Some studies on the health effects of PCBs are confounded by exposure to other halogenated environmental contaminants and by impurities in the PCBs, notably chlorinated dibenzofurans. This CICAD deals only minimally with the toxicity of contaminants that result from either the manufacturing process or the heating of PCBs (e.g., PCDDs, PCDFs, or even other persistent organic pollutants); however, studies dealing with the Yusho and Yu-Cheng contaminated cooking oil accidents are briefly summarized in this document.

In studies on humans exposed to PCBs, effects on sperm motility, fetal growth rate (lower birth weight, smaller head circumference) and development (shorter gestational age, neuromuscular immaturity), and neurological functions of the offspring (impaired autonomic function, increased number of abnormally weak reflexes, reduced memory capacity, lower IQ scores, and attention deficit) have been observed. Some of the neurological deficiencies at early ages may disappear later during childhood.

Epidemiological studies suggest exposure-related increases in cancers of the digestive system, especially liver cancer, and malignant melanoma. However, the limitations of exposure information, the inconsistency of the results, and, in some cases, the presence of confounding exposures preclude a clear identification of an exposure–response relationship.

No increase in the incidence of respiratory tract infections during the first 18 months of life was observed, but changes in the relative amounts of different circulating lymphocyte types were observed among children born to PCB-exposed mothers. Decreased numbers of natural killer cells have been observed in consumers of PCB-contaminated fish. The prevalence of recurrent middle-ear infections and chicken pox was related to plasma PCB concentrations in 3.5-year-old children.

Adverse health effects were observed in rats, mice, monkeys, and other mammalian species. Effects were seen in most animal health end-points, such as immunological, developmental, reproductive, hepatic, and body weight. Several studies consistently report an increase in liver cancer incidence among rodents exposed to different PCBs. The severity of the health effects depended on dose, species, PCB mixture, duration and timing of the exposure, and other factors.

Limited studies indicate that PCBs are not genotoxic by direct mechanisms.

Secondary challenge with sheep red blood cells after exposing monkeys to PCBs for 55 months showed decreasing trends in the IgM and IgG anamnestic responses, with IgM significantly lower than in controls for all doses. Based on a lowest-observed-adverse-effect level (LOAEL) of 5 μg/kg body weight per day for several end-points, a tolerable intake of 0.02 μg/kg body weight per day for an Aroclor 1254 mixture was derived, using an overall uncertainty factor of 300 (10 for use of a LOAEL rather than a no-observed-adverse-effect level [NOAEL], 3 for interspecies variation, and 10 for intraspecies variation).

PCBs are a class of chemical compounds in which chlorine atoms replace some or all of the hydrogen atoms on a biphenyl molecule. The general chemical structure of chlorinated biphenyls is shown below in Figure 1.

Fig. 1: Biphenyl molecule with the numbering system. In PCBs, some or all of the 10 hydrogens
(attached to carbon atoms numbered 2–6 and 2'–6') are substituted with chlorines.

PCBs were manufactured and sold as mixtures with a variety of trade names, including Aroclor, Pyranol, Pyroclor (USA), Phenochlor, Pyralene (France), Clopehn, Elaol (Germany), Kanechlor, Santotherm (Japan), Fenchlor, Apirolio (Italy), and Sovol (USSR).

Two different but correlated nomenclature systems are currently used. The IUPAC name (according to IUPAC rules A-52.3 and A-52.4) identifies the numbered carbons to which chlorines are attached and lists the numbers sequentially (e.g., the PCB congener with chlorines on carbons 2, 3, 4, and 3' is identified as 233'4); a variant of that system lists the chlorinated ring positions separately, sometimes eliminating the prime symbols for clarity and ease of typing (e.g., 234-3' or 234-3). A second widely used system was developed by Ballschmiter & Zell (1980) as a way to simplify referring to specific congeners. It correlates the structural arrangement of the PCB congeners in an ascending order of number of chlorine substitutions within each sequential homologue. An unprimed number is considered lower (higher priority) than the same number when primed. This results in the congeners being numbered from PCB 1 through PCB 209. Original typographical errors in the Ballschmiter & Zell (1980) numbering system have subsequently been resolved (i.e., old PCB numbers 107, 108, 109, 199, 200, and 201 are now numbered 109, 107, 108, 200, 201, and 199, respectively). Table 1 shows the relationship between the IUPAC and revised PCB numbering systems. Some of the congeners that are either prevalent in commercial PCB products or considered by some researchers to be more toxic than average PCBs are shown in the tables cited later in this section. The primary focus of this CICAD is on commercial mixtures of PCBs and mixtures generated from them in the environment and food-chain.

Table 1: PCB nomenclature conversion table.^a

As evidenced in Table 1, 209 chlorinated compounds, called congeners, are possible. PCBs can also be categorized by degree and location of chlorination. The term "homologue" is used for all compounds with the same number of chlorines (e.g., congeners with three chlorines attached are termed trichlorobiphenyls). PCBs of a given homologue with different substitution patterns are called isomers. There are, for example, 12 isomers in the dichlorobiphenyl homologue.

The benzene rings can rotate around the bond connecting them, but the rings are forced towards either the same plane (called planar or coplanar) or perpendicular planes (termed non-planar) by the electrostatic repulsion of the highly electronegative chlorine atoms. The degree to which the rings can twist beyond these two extremes is a function of steric hindrance produced by chlorine atoms in different positions on the two rings. A non-planar orientation is produced by multiple substitutions in the ortho positions (2, 2', 6, and 6'), and an increase from two to four substitutions results in a progressively stronger rotational hindrance. Conversely, some mono-ortho-substituted and all non-ortho-substituted PCBs are considered to be planar, otherwise called coplanar or mono-ortho coplanar, implying that the rings of some congeners can twist but not turn completely. Additionally, solely ortho-substituted congeners, on one or both rings, may be polar molecules with an ability to form hydrogen bonds and thus may be more water soluble. Meta- and para-saturated congeners would be more non-polar and so more lipid-soluble. The congeners considered to be most toxic, based on combined health effects considerations, are coplanar.

An important property of PCBs is their general inertness. PCBs resist both acids and alkalis and have thermal stability, making them useful in a wide variety of applications, including dielectric fluids in transformers and capacitors, heat transfer fluids, and lubricants (Afghan & Chau, 1989).

At high temperatures, PCBs are combustible, and the products of combustion may be more hazardous than the original material. Combustion by-products include hydrogen chloride and PCDFs. Combustion of technical-grade materials containing PCBs and chlorobenzenes (such as some dielectric fluids) may also produce PCDDs (IPCS, 1993; ATSDR, 2000). PCDFs are also produced during commercial production and handling of PCBs. The amount of PCDFs formed depends upon manufacturing conditions. The impurities 2,3,7,8-tetrachlorodibenzofuran and 2,3,4,7,8-pentachlorodibenzofuran were found at concentrations of 0.33 and 0.83 mg/kg, respectively, in Aroclor 1248; and at 0.11 and 0.12 mg/kg, respectively, in Aroclor 1254. Concentrations of PCDFs in commercial PCB mixtures, including Clophen A-60, Phenoclor DP-6, and Kanechlor 400, have been reported.

Physical properties such as solubility, vapour pressure, and Henry’s law constant have been reported for individual congeners. Experimentally determined octanol/water partition coefficients (K_ow values) for 19 congeners and an estimation method for the determination of log K_ow values of other PCB congeners are also available. The congeners reported are important due to their toxicity or because they occur in high concentrations in the environment. A comprehensive database of chemical and physical data exists (Syracuse, 2000). Table 2 contains such values for the most toxic and the most environmentally prevalent congeners. Tables 3 and 4 summarize the compositions of common Aroclor mixtures by congener prevalence and congener toxicity, while Table 5 categorizes the Aroclors by homologue. Generally, PCBs are relatively insoluble in water, with the highest solubilities among the ortho-chlorinated congeners (5 mg/litre for PCB 1), which may be due to hydrogen bonding associated with the more polar character of these molecules. Solubility decreases rapidly in ortho-vacant congeners, especially as the para positions are filled, which may result in greater and more uniform perimeter electronegativity and interference with hydrogen bonding. PCBs are freely soluble in non-polar organic solvents and biological lipids (US EPA, 1980), and the shift from water to lipid solubility is shown in Table 2 as an increasing K_ow with increased chlorination. PCBs, especially the more chlorinated congeners, are also relatively non-volatile, with partial pressures and Henry’s law constants that tend to decrease with increased chlorination, especially for meta- and para-saturated congeners.

Table 2: Physical and chemical properties of some of the most toxic and/or environmentally prevalent PCB congeners.^a,b

Table 3: Most prevalent congeners (mol%) of common commercial PCB products.^a

^a Values less than approximately 4% are not included.

Table 4: Percent composition for some of the most toxic congeners in commercial Aroclors (mol%).^a

Table 5: Estimated homologue composition (%) of different Aroclors.

The quantification of PCBs in biological samples usually consists of three distinct steps: extraction of PCBs from the sample matrix by a solvent or a combination of solvents, cleanup of PCBs (i.e., removal of impurities) on single or multiple columns, and quantification by gas chromatography (GC) with a suitable detector. Authors may report PCB concentrations as Aroclors, as sum of homologues, or as individual congeners.

PCBs are extracted from blood or serum using hexane, benzene, or mixed solvents, such as hexane/ethyl ether. A variety of adsorbents may be used for cleanup and/or fractionation of extracts: deactivated silica gel (Burse et al., 1989), Florisil, alumina (Koopman-Esseboom et al., 1994), or multiple columns. Supercritical fluid extraction (SFE) has also been used to extract PCBs from adipose tissue samples (Djordjevic et al., 1994). GC coupled with an electron capture detector (GC/ECD) is used most often to determine PCBs (Burse et al., 1989; ATSDR, 2000), but confirmation by mass spectrometry (MS) is recommended when multiple individual congener measurements are required. Detection limits for individual Aroclors are in the low- to sub-microgram per litre range, and recoveries, where reported, range from 80% to 96% (Koopman-Esseboom et al., 1994; ATSDR, 2000). The accuracy and precision of the results of PCB analysis in serum, using a packed column GC/ECD method, were examined in a collaborative study (Burse et al., 1989). Capillary or high-resolution gas chromatography (HRGC) has made it possible to achieve lower detection limits and better separation of PCB congeners for quantification (Mullin et al., 1984). Although complete separation with a single column has not yet been achieved, advances have been made in analysing for specific coplanar PCBs (77, 126, 169), which some consider to be the most toxic congeners. Equipment can be calibrated to yield results by congener; however, because of the time required to analyse all congeners, results may be reported only for selected ones or for homologues, using the areas or heights of selected peaks or ranges to estimate the concentration in a sample. This tends to reduce analytical costs but can also complicate study intercomparisons. Detection limits have been reported as 144 ng/g for adipose tissue, 2 ng/g for blood, 0.01 ng/g for plasma, and 1–2.5 ng/ml for serum.

Traditional PCB analytical methods quantify PCBs as Aroclor mixtures, and one assumption has been that the original congener formulation is retained in the environment. The validity of this assumption suffers as individual congeners undergo different physical, chemical, and biological interactions with the environment that alter the congener mixture relative to the original formulation. Analysis by Aroclor also has the disadvantage of being insensitive to the 4-to-6-chlorine coplanar congeners that are of highest potential biological significance for some health effects end-points. An adjusted approach is to analyse the isomer classes (C1–C10), nine of which are addressed in Table 5.

The most appropriate approach is to analyse for individual congeners. Results may then be adjusted to compare with older studies, where mixtures were assessed from multiple peaks. Summing of selected congeners is also often done; sums can be weighted for biological importance, and summing systems form the basis for long-term monitoring programmes such as the National Oceanic and Atmospheric Administration Mussel Watch Program (NOAA, 1989). McFarland & Clarke (1989) recommended including priority congeners 49, 77, 87, 101, 105, 118, 126, 128, 138, 153, 156, 158, 169, 170, 180, 183, and 184. These include congeners with greatest environmental importance based on potential toxicity, frequency of occurrence in environmental samples, and relative abundance in animal tissue. Congener analysis can be performed using high-resolution capillary gas chromatograph columns and electron capture detectors (HRGC/ECD) or mass spectrometric (HRGC/MS) techniques; the latter has individual congener detection levels approaching 0.01 ng/litre in human serum. Ballschmiter & Zell (1980) and Safe et al. (1985b) reported a congener-specific analysis of a commercial PCB preparation and the PCB composition of a human milk sample.

Studies have described the analysis of non-ortho coplanar and mono-ortho coplanar PCBs in breast milk and coplanar PCBs in serum and adipose tissue. Determination of these congeners, which are the most toxic, is useful in assessing the toxic potential of breast milk for infants and of dietary intake for adults.

Sampling method variables may also greatly influence results. Random error, interlaboratory variations in procedure, and methods used for reporting data may have considerable impact on the reported PCB levels in human tissues. Since there is no standard method or approach for analysing PCBs in biological samples, caution is appropriate when comparing exposure estimates or health effects studies reported by different investigators.

Methods are available for measuring concentrations of PCBs in fish and animal tissues. Tissues are homogenized, then dried by blending with sodium sulfate prior to Soxhlet or column extraction. Few methods are available to determine PCBs in foods. A minimum of performance data has been reported as well. A method is available to determine Aroclors in poultry fat, fish, and dairy products. Congener-specific analytical methods for coplanar PCBs 77, 126, and 169 are amenable to these media.

Air samples are usually collected by pumping air through a sampler containing a glass fibre filter and adsorbent trap to separate the particle-bound and vapour-phase fractions. Adsorbents used most often include Florisil and polyurethane foam. Florisil traps are solvent desorbed, and XAD-2 traps are Soxhlet extracted. PCBs are determined by GC/ECD or HRGC/MS. Detection limits of low nanograms per cubic metre for individual Aroclors to low picograms per cubic metre for individual congeners have been reported. Recovery, where reported, is good, greater than 80%.

Drinking-water samples are typically extracted with solvent prior to analysis by GC/ECD or HRGC/ECD. Composite PCB detection limits are in the sub-microgram per litre range, and recovery is greater than 80%. Method 508A of the US Environmental Protection Agency (EPA), which converts all the PCBs to decachlorobiphenyl, is a screening method for quantifying total PCBs. The method is likely to show interference because of perchlorination of biphenyl or related compounds and because the method cannot quantify either the individual commercial Aroclors in a PCB mixture or the individual congeners present.

Soil, sediment, and solid waste samples are usually Soxhlet extracted. Ultrasonic extraction with various solvent combinations and SFE are also used. Analyses for screening PCB contamination in soils, using the enzyme-linked immunosorbent assay, are commercially available. These methods are inexpensive and provide fast turnaround time.

A number of Standard Reference Materials (SRMs) with certified PCB congener concentrations are available from the US National Institute of Standards and Technology. These include SRM 1588 for cod liver oil, SRM 1939 for river sediment, SRM 1941 for marine sediment, and SRM 1974 for mussel tissue.

PCB production started in the late 1920s. Since 1929, about 2 × 10⁹ kg of PCBs have been produced commercially, of which about 2 × 10⁸ kg remain in mobile environmental reservoirs. PCB pollution may occur during the incineration of municipal waste. PCB concentrations of 0.01–1.5 mg/kg were detected in fly ash from five municipal incinerators operating under different technological and working conditions. Stack effluents from several municipal refuse and sewage incinerators in the midwestern USA contained PCB concentrations of 0.3–3.0 μg/m³. The total PCB concentration measured in the flue gas effluent from a municipal waste incinerator in Ohio, USA, was 0.26 μg/m³. PCB levels of 2–10 ng/m³ were detected in effluents from coal and refuse combustion in Ames, Iowa, USA (US EPA, 1988a). An additional source of PCB pollution is volatilization from landfills containing transformers, capacitors, and other PCB waste and from contaminated bodies of water, such as the Great Lakes in North America. Because of possible health implications and environmental impacts, the use and production of PCBs are severely restricted or banned in many countries. Sweden restricted their use and production in 1972, the USA in 1977, Norway in 1980, Finland in 1985, and Denmark in 1986.

Values for estimated Henry’s law constants for Aroclors, ranging from 29 to 47 Pa·m³/mol, indicate that vaporization can be an important environmental transport process for PCBs dissolved by surface waters (Thomas, 1982) when the concentrations in silt force the water levels to remain elevated and evaporation represents a significant portion of the overall water loss. The volatility and solubility differences among the various congeners can be expected to cause redistribution in both surface water and bottom sediment; this emphasizes the need for congener-specific analysis of environmental and biological samples. A study of Lake Michigan in North America indicates that vaporization may be the major process for the removal of PCBs from lakes (Swackhamer & Armstrong, 1986). PCBs may vaporize even more significantly from dam spillways and outfalls, waterfalls, and other waterways with markedly higher aeration rates (McLachlan et al., 1990). Nonetheless, adsorption to sediment significantly decreases the rate of vaporization of highly chlorinated Aroclors from the aquatic phase (Lee et al., 1979; US EPA, 1985a,b).

In water, adsorption to sediments and suspended particulates is also a major process that partitions PCBs from the water to the solid phase. A study of the Saginaw River in Michigan, USA (Verbrugge et al., 1995), reports that the ratio of total PCBs bound to suspended particulates relative to dissolved PCBs was 2:1 and that this ratio remained fairly constant for river discharges of less than 400 m³/s. The adsorption usually increases as organic matter, clay, and micro-particle contents in the water increase (US EPA, 1980).

Adsorption and subsequent sedimentation may immobilize PCBs for a long time in aquatic systems. Redissolution of PCBs into the water column and vaporization from the water surface into the air occur in the environment. Therefore, the substantial quantities of PCBs contained in aquatic sediments can act as an environmental reservoir from which PCBs may be released slowly over a long period. The rate of redissolution of PCBs from sediment to water and of evaporation from water to air is always greater in summer than in winter, because these parameters increase with temperature (Larsson & Soedergren, 1987).

PCBs in the atmosphere are physically removed by wet and dry deposition (Eisenreich et al., 1981; Leister & Baker, 1994). Dry deposition of PCBs occurs from gravitational settling of particulates and from impaction of vapour-phase PCBs on land or aquatic surfaces. Wet deposition of PCBs occurs through rain, snow, and fog (Hart et al., 1993). The atmospheric wet and dry deposition fluxes of PCBs over Chesapeake Bay in Maryland, USA, in 1990–1991 were 1.9 μg/m³ per year and 1.4 μg/m³ per year, respectively (Leister & Baker, 1994). Thus, wet depositional flux in the bay represented 58% of the overall flux. This is a function of the local precipitation pattern; overall, more wet deposition results from particle washout (99%) than from vapour washout (1%) (Atlas & Giam, 1987).

The measured atmospheric concentration of PCBs near a Superfund⁶ hazardous waste site was higher than atmospheric concentrations monitored 15 km from the site (Hermanson & Hites, 1989), suggesting that PCBs may be transported from soil to air through vaporization and subsequently diluted downwind. The vaporization rate was greater from soil with low organic carbon content, because of the weaker adsorption of PCBs (Shen & Tofflemire, 1980). Vaporization rates are also greater in moist soils because of the co-evaporation of PCBs with water. Storm water runoff can also transport PCBs through soil erosion.

Because of their large and mobile biomass, position in the food-chain, and biotransformation potential, insects may significantly contribute to the global transport and transformation of PCBs (Saghir et al., 1994).

The bioconcentration factor (BCF, or ratio of the concentration of PCBs in the organism to the concentration of PCBs in water) values of PCBs are expected to increase with an increase in chlorine substitution and a decrease in water solubility (Zhang et al., 1983). However, the maximum bioaccumulation was observed for hexachlorobiphenyls and not hepta- or octachlorophenyls (Porte & Albaiges, 1993; Bremle et al., 1995). The lower BCF values for the latter two classes of higher chlorinated compounds may be due to lower uptake rates.

The elimination of PCBs from aquatic organisms is both species- and congener-specific. Generally, congeners containing two vicinal hydrogen atoms at the meta and para positions in at least one aromatic ring are easily metabolized (Pruell et al., 1993). The biotransformation of PCBs in vertebrates is mediated by the cytochrome P450-dependent mixed-function oxygenase (MFO) (Safe et al., 1985a). There is evidence that different cytochrome P450 enzymes metabolize specific PCB congeners. In rats, the di-ortho PCBs are preferably metabolized by the CYP2B family, while the CYP1A enzymes preferably metabolize the non-ortho PCBs (Kaminsky et al., 1981). Cytochrome P450-dependent MFO activities are species-dependent, and MFO activities are generally much lower in mussels than in crabs and fish (Porte & Albaiges, 1993). The PCB congeners 110, 138, 149, 153, and 187 are most recalcitrant in mussels. The most stable PCB congeners are 138, 153, 170, and 180 in crabs; 138, 153, 170, 180, and 187 in mullet; and 84, 110, 118, and 138 in tuna (Porte & Albaiges, 1993).

Bioaccumulation in tuna of PCB congeners 84 and 110, which are rapidly metabolized by birds and mammals (Hansen, 1987), is unusual and may be related to the seasonal shallow surface feeding habit of tuna. The 2,3,6 congeners (including 149) are also more volatile (Mullin et al., 1984) and are likely found at higher concentrations near the air and water interface. Bioaccumulation of PCBs in aquatic animals may also depend on the water zone in which the animals predominantly reside and feed. When airborne PCBs are deposited onto the surface of water, they become enriched in the surface strata. The PCB levels are 500 times higher in the surface strata than in deeper water. Consequently, bioaccumulation by fish is several times higher in this zone (Soedergren et al., 1990). Because the concentration of PCBs in sediments is several times higher than in water, the bioaccumulation of PCBs in bottom-feeding species is also high. Loss of accumulated PCB residues from the tissues is slow, because PCBs tend to remain stored in lipids. Therefore, greater bioaccumulation occurs in the fatty tissues than in the muscles or whole body of aquatic organisms (US EPA, 1980). However, studies of fish indicate that stored PCBs may become more mobilized from lipids of organs that contain higher polar lipid components, such as phospholipids (Boon et al., 1984).

PCBs generally biomagnify within aquatic food-chains, as indicated by the PCB concentrations detected in higher trophic levels of aquatic organisms (LeBlanc, 1995; Wilson et al., 1995). This biomagnification is evident in shellfish that accumulate PCBs from consumption of phytoplankton and zooplankton (Secor et al., 1993) and in marine mammals (seals, dolphins, and whales) that accumulate PCBs from consumption of plankton and fish (Andersson et al., 1988; Kuehl et al., 1994; Kuehl & Haebler, 1995; Lake et al., 1995; Salata et al., 1995). Food-chain biomagnification also occurs in several species of fish-consuming birds (Mackay, 1989). Concentrations of PCBs in common (Sterna hirundo) and Forster’s terns (S. forsteri) (which are primarily piscivores) are higher than concentrations in tree swallows (Tachycineta bicolor) and red-winged blackbirds (Agelaius phoeniceus) (which are insectivores) (Ankley et al., 1993), which shows evidence of biomagnification of PCBs in the aquatic food-chain. The biomagnification of PCBs in the aquatic food-chain is also congener-specific (Koslowski et al., 1994). For example, although the concentrations of congener 138 increased from plankton (1 μg/kg) to piscivores (1388 μg/kg in silver bass [Morone chrysops] muscle tissue) to herring gulls (Larus argentatus) (30 063 μg/kg), the three toxic congeners 77, 126, and 169 showed no obvious biomagnification in these species. The lack of biomagnification in congener 77 was attributed to its rapid elimination by aquatic species (Koslowski et al., 1994). These samples were collected during the summer of 1991.

Biotransfer factors for organic chemicals in beef and milk are directly proportional to their K_ow values. The biotransfer factors of Aroclor 1254 (concentration in food [mg/kg]/daily intake of Aroclor [mg/day]) for beef and milk are estimated at 0.052 and 0.011 kg/day, respectively, using the estimation procedures of Travis & Arms (1988). Based on summary data from Canada, the estimated mean BCF value (the ratio of the concentration of PCBs in tissues to the concentration of PCBs in the diet) for PCBs in human fat is 128 on a wet weight basis and 192 on a lipid weight basis (Geyer et al., 1986). The biomagnification of PCBs in the terrestrial food-chain occurs through accumulation of PCBs from soil/plant to earthworm to bird (Hebert et al., 1994) and from oak leaves to caterpillars to birds (Winter & Streit, 1992). While total PCBs were not detected in soil or plants, the ranges in concentrations (wet weight) observed were 14.8–18.6 μg/kg in earthworms; not detected to 208.8 μg/kg in mammals; 39.2–68.3 μg/kg in starlings; 71.5–157.2 μg/kg in robins; and 56.0–219.9 μg/kg in kestrels (Hebert et al., 1994). In addition, the authors reported that the sum of PCBs in pooled egg samples ranged from 0.066 to 0.477 mg/kg for kestrels and from 0.032 to 0.061 mg/kg for bluebirds. A concentration of 5.298 mg/kg was detected in a single pooled egg sample from herring gulls, a piscivorous species. Juvenile great tits (Parus major) received PCBs from the mother bird through egg transfer and from caterpillars, their primary food source (Winter & Streit, 1992).

The ability of PCBs to be degraded or transformed in the environment depends on the degree of chlorination of the biphenyl molecule and on the chlorination pattern (Callahan et al., 1979; Leifer et al., 1983; US EPA, 1988a). Generally, the persistence of PCB congeners increases as the degree of chlorination and structural uniformity increase. Adjacent unchlorinated carbons allow the formation of arene oxide intermediates and thus facilitate metabolism. Kubatova et al. (1998) examined the biodegradation of PCB congener 11 (3,3'-dichlorobiphenyl) in the soil. PCB 11, labelled with ¹⁴C, was chosen as a low chlorinated coplanar congener and assumed to be readily degraded by Pseudomonas species and the oyster mushroom (Pleurotus ostreatus). After a 2-month incubation, results showed that the mineralization of PCB 11 was <0.4%, volatilization was <3.1%, and 30% of the radioactivity was irreversibly bound to the soil matrix. The major biodegradation product was 3-chlorobenzoic acid. The concentrations of the coplanar congeners were significantly lowered by reductive dechlorination and by anaerobic bacteria (Quensen et al., 1998).

5.2.1 Air

In the atmosphere, the vapour-phase reaction of PCBs with hydroxyl radicals (which are photochemically formed by sunlight) may be the dominant transformation process. The estimated tropospheric half-times for this reaction increase as the number of chlorine substitutions increases. The half-times are 3.5–7.6 days for monochlorobiphenyl, 5.5–11.8 days for dichlorobiphenyl, 9.7–20.8 days for trichlorobiphenyl, 17.3–41.6 days for tetrachlorobiphenyl, and 41.6–83.2 days for pentachlorobiphenyl (Atkinson, 1987). Photochemical studies conducted under simulated and natural sunlight with a number of chlorobiphenyl congeners and commercial PCB mixtures in aqueous suspension, thin film, or vapour state resulted in several degradative reactions that produced dechlorination, polymerization, and polar (hydroxy- and carboxy-) products (Hutzinger et al., 1972).

5.2.2 Water

In water, transformation processes such as hydrolysis and oxidation do not significantly degrade PCBs (Callahan et al., 1979). Photolysis appears to be the only viable chemical degradation process in water. PCBs containing up to six chlorine substitutions do not significantly absorb sunlight, and the estimated photolysis half-times of mono- through tetrachlorobiphenyls with summer sunlight at a shallow water depth (<0.5 m) range from 17 to 210 days. Photolysis rates with sunlight are even slower during winter. Nonetheless, as the number of chlorine substitutions increases, the light absorption band shifts towards longer wavelengths, and the photolysis rate for hepta- through decachlorobiphenyls increases.

Biodegradation of PCBs in water, although slow, can occur under both aerobic and anaerobic conditions. However, biodegradation is probably more significant in soil and sediment than in water, given the higher numbers of microorganisms. The use of adapted (pre-exposed) microbial populations and the addition of amenable substrates for co-metabolic and co-oxidative biotransformation can enhance the biotransformation and biodegradation of PCBs.

5.2.3 Sediment and soil

Biodegradation occurs under both aerobic and anaerobic conditions and is the major degradation process for PCBs in soil and sediment. No abiotic process is known to significantly degrade PCBs in sediment and soil; however, photolysis of PCBs on surface soil may occur. Higson (1992) and Robinson & Lenn (1994) provide a review of biodegradation of PCBs in soil and sediment. Experiments with pure and mixed cultures of microorganisms show that some congeners of PCBs, usually containing six or fewer chlorine substituents, biodegrade under aerobic conditions (Leifer et al., 1983; US EPA, 1988a; Sugiura, 1992; Thomas et al., 1992; Dowling et al., 1993; Fava et al., 1993; Gibson et al., 1993). Biodegradation rates are highly variable because they depend on several factors, including the amount and location of chlorination, PCB concentration, type of microbial population, available nutrients, and temperature. The most common process for the aerobic degradation of PCBs by bacterial cultures proceeds by two distinct steps, one involving bioconversion of PCBs to chlorinated benzoic acids and the other involving mineralization of chlorobenzoates to carbon dioxide and inorganic chlorides (Thomas et al., 1992; Robinson & Lenn, 1994). Complete mineralization of biodegradable PCBs requires the presence of two clusters of genes for the two-step bioconversion process (Sondossi et al., 1992); therefore, complete degradation requires mixed microbial cultures (Afghan & Chau, 1989).

Reliable evaluation of the potential for human exposure to PCBs depends partly on the reliability of supporting analytical data from environmental samples and biological specimens. With respect to PCB analysis, comparisons among various studies are complicated by the fact that authors may report PCB concentrations as Aroclors, homologues, or congeners.

Table 6 lists PCB levels relevant to human exposure in air and water in various countries. The mean atmospheric concentration of PCBs in urban areas in the USA was 5 ng/m³ (range 1–10 ng/m³) (Eisenreich et al., 1992). The mean atmospheric concentrations of PCBs in two rural areas (rural Ontario, Canada, and Adirondack, New York, USA) were 0.2 and 0.95 ng/m³, respectively (Knap & Binkley, 1991; Hoff et al., 1992). These values are within the range for continental areas (0.2–1.5 ng/m³) given by Eisenreich et al. (1992). For two remote areas (the Arctic and the Antarctic), the mean PCB value was 0.2 ng/m³ (range 0.02–0.5 ng/m³) (Tanabe et al., 1983; Baker & Eisenreich, 1990). The air concentration in the eastern Arctic in 1996 was 0.074 ng/m³ (Harner et al., 1998).

Table 6: PCB levels in air and water in various countries.^a

^a Adapted from IPCS (1993).

The mean atmospheric PCB concentration in marine and coastal areas was 0.1 ng/m³ (range 0.01–0.7 ng/m³). Over the Great Lakes in North America, the mean concentration was 1.0 ng/m³ (range 0.2–4.0 ng/m³). These samples were collected in the late 1980s to early 1990s (Eisenreich et al., 1992).

Since the early 1980s, PCB concentrations in the air have shown a slightly decreasing trend for urban, rural, and marine/coastal areas (Eisenreich et al., 1981). Over the same approximate period (late 1970s to early 1980s compared with late 1980s to early 1990s), rainwater PCB concentrations from continental areas significantly declined by 4- to 5-fold, with values decreasing from 20 to 5 ng/litre in rural areas and from 50 to 10 ng/litre in urban areas.

When indoor air in a number of laboratories, offices, and homes was monitored for various Aroclors, the "normal" indoor air concentrations of PCBs were at least 1 order of magnitude higher than those in the surrounding ambient outdoor atmosphere (MacLeod, 1981). For example, average PCB levels were 100 ng/m³ inside an industrial research building and 210 ng/m³ inside the laboratories, compared with <20 ng/m³ in air outside the facility. The mean PCB indoor air concentration in one home was 310 ng/m³, while the average air concentration outside the home on the same day was 4 ng/m³. Certain electrical appliances and devices (e.g., fluorescent lighting ballasts) and building materials (elastic sealant), which have PCB-containing components, may emit PCBs into the indoor air, thereby elevating indoor PCB levels significantly above outdoor background levels (Balfanz et al., 1993).

The average PCB concentration (Aroclors 1242 and 1260) in emissions from gas vents at a hazardous waste landfill in North Carolina, USA, was 126 000 ng/m³ (Lewis et al., 1985). The maximum total PCB concentration detected in air samples collected at Raquette Point within the Mohawk Nation Reservation at Akwesasne, New York, USA (adjacent to a Superfund site), was 50 ng/m³ (ATSDR, 1995).

PCB concentrations reported for seawater from various oceans include 0.04–0.59 ng/litre in the North Pacific, 0.035–0.069 ng/litre in the Antarctic, and 0.02–0.20 ng/litre in the North Atlantic (Giam et al., 1978; Tanabe et al., 1983, 1984). PCB levels were several orders of magnitude higher in sea surface microlayer samples taken from industrial areas than in those taken from sites farther offshore (Cross et al., 1987). This indicates that the PCB contribution from anthropogenic sources is higher in nearshore samples than in offshore samples. PCB concentrations of 0.3–3 ng/litre, which are higher than the seawater concentrations reported above, have been detected in surface seawater from the North Sea (Boon & Duinker, 1986). Analysis of water from eight sites in Galveston Bay, a highly industrialized area in Texas, USA, showed an average PCB concentration of 3.1 ng/litre between 1978 and 1979 (Murray et al., 1981). In 1993, the total dissolved plus particulate PCB concentration in Great Lakes water ranged, in increasing order, from 0.07 to 0.10 mg/litre, from 0.12 to 0.16 mg/litre, from 0.17 to 0.27 mg/litre, from 0.19 to 0.25 mg/litre, and from 0.2 to 1.6 mg/litre for Lakes Superior, Huron, Michigan, Ontario, and Erie, respectively (Anderson et al., 1999).

The historic profiles of PCB concentrations in sediments of the lower Passaic River, New Jersey, USA, were studied by determining the concentrations at different depths. The authors concluded that total PCB sediment concentrations peaked at 4.7 mg/kg dry weight in the 1970s and then decreased to 1.1 mg/kg dry weight in the 1990s (Wenning et al., 1994). A similar study of dated sediments from the Newark Bay Estuary, New Jersey (including the Passaic River), also reported that the highest concentration of PCBs occurred in buried sediments from the Passaic River and Newark Bay, corresponding to historic deposition during the 1960s and 1970s, the peak manufacturing period for Aroclors (Iannuzzi et al., 1995).

Yao et al. (2000) analysed sediment core samples attributable to the period from 1935 to 1993 in Tokyo Bay, Japan. Coplanar PCB concentrations peaked from 1967 to 1972 (reflecting the historical production and use patterns), decreased rapidly from 1972 to 1977, following the ban on production and use of PCBs, and then slowly levelled off to about one-third of the peak level. Kang et al. (2000) reported a decline in PCB levels in fish from 1953 to 1975 that was similar to the decline in PCB levels in sediment.

An 80% decrease (from 9.08 μg/g wet weight in 1976 to 1.72 μg/g in 1994) in PCB concentration in trouts from Lake Ontario was reported by Huestis et al. (1996); a similar decrease was noted in the two species of trout tested (lake trout [Salvelinus namaycush] and rainbow trout [Oncorhynchus mykiss]) from both Lake Ontario and Lake Huron. In several fish species from the Great Lakes, the PCB concentrations in samples collected in the 1990s were generally below 1 µg/g wet weight (ATSDR, 2000).

The general population is exposed to PCBs via air, drinking-water, and food.

Typically, outdoor air in urban areas contains an average PCB concentration of 5 ng/m³ (Eisenreich et al., 1992). As the average adult male inhales 23 m³/day (IPCS, 1994), the average daily exposure through inhalation is approximately 100 ng. However, the concentration of PCBs can be at least an order of magnitude higher in indoor air than in outdoor air. PCB concentrations in the workplace air of unspecified PCB disposal facilities in the USA ranged from 850 to 40 000 ng/m³. In 95 of the 96 air samples collected for analysis from such facilities, PCB concentrations exceeded 1000 ng/m³ (Bryant et al., 1989).

Based on a US EPA (1988a) study, PCB levels in drinking-water were below 100 ng/litre, and not all water samples had detectable PCB levels. Using a conservative adult consumption rate of 2 litres/day, the general US population is exposed to <200 ng PCBs/day from drinking-water.

In a Canadian drinking-water survey conducted in 1985–1988 (O’Neill et al., 1992), PCBs were detected in 1 out of the 280 municipal drinking-water samples, at a level of 6 ng/litre.

The primary exposure pathway appears to be through consumption of contaminated foods, particularly meat, fish, and poultry (ATSDR, 2000). In the USA, the dietary intake of PCBs for adults continually decreased from 1978 until 1986–1991 (Table 7). The mean daily intake was <0.001 µg/kg body weight per day for individuals of ages 6 months and 25–30 years and 0.002 µg/kg body weight per day for 2-year-olds between 1986 and 1991 (based on average total diet composition) (Gunderson, 1995). In a study on dietary intake during 1991–1997, the decreasing trend did not continue, and the dietary exposure was 3–5 ng/kg body weight per day for adults and 2–12 ng/kg body weight per day for children of different ages (based on calculated intakes from 265 different food items) (P.M. Bolger, personal communication, 1999, as cited in ATSDR, 2000).

Table 7: Estimated daily dietary intake of PCBs in the USA.^a