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A Risk-Management Strategy for PCB-Contaminated Sediments 2 PCBs in the Environment This chapter provides an overview of PCBs in the environment as a background to understanding their history of use, sources of input to the environment, distribution in the environment, and their human health and ecological effects. Because PCBs are such complex chemicals, knowledge of their chemical and physical properties is needed to understand their transport, fate, and toxicity. Considerable new information has become available in the past two decades and new information from field and laboratory studies is published regularly. DEFINING PCBs Polychlorinated biphenyls, or PCBs, as they are commonly called, have been used industrially since 1929 (Jensen 1972), and are entirely of anthropogenic origin. The backbone of the chemical structure is a biphenyl, consisting of two hexagonal “rings” of carbon atoms connected by carbon-carbon bonds. The specific manner by which the carbon atoms share electrons forming the hexagonal rings leads to the biphenyl being an “aromatic” compound. Polychlorinated biphenyls have between 1 and 10 chlorine atoms substituting for hydrogen atoms on the biphenyl rings (Figure 2–1). The various number and positions of the chlorine atoms on the biphenyl molecule result in up to 209 possible chemical structures designated as congeners in the scientific literature. PCBs are subdivided into groups based on the degree of chlorination or number of chlorine atoms per biphenyl molecule (e.g., trichlorobi-
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A Risk-Management Strategy for PCB-Contaminated Sediments FIGURE 2–1 Synthesis of PCBs (e.g., 2,3′,4,5,5′-pentachlorobiphenyl) by direct chlorination of biphenyl. phenyls (three chlorines) and tetrachlorobiphenyls (four chlorines)). The PCBs within a series of structures of specific chlorine content are known as homologues (i.e., the mono-, di-, tri-, tetra-, penta-, hexa-, hepta-, octa-, nona-, and decachlorobiphenyl homologues). Within a homologue group (e.g., the trichlorobiphenyls), the individual chlorobiphenyl molecules are isomers of each other, meaning that they each have the same number of chlorine atoms, but these chlorine atoms are arranged at different positions on the biphenyl rings. Examples of chemical structures of PCBs are provided in Figures 2–2 and 2–3. A complete list of congeners is in Appendix H. Since the chlorine atom is part of the group of elements known as halogens (others are fluorine, bromine, and iodine), polychlorinated biphenyls are part of a larger group of chemicals known as halogenated aromatic compounds. Industrial PCBs were complex mixtures composed of up to 50 or 60 congeners (or individual chlorobiphenyls). The composition of the PCB mixture was governed by the reaction conditions and the reaction properties by which they were manufactured. These conditions and properties favor the production of specific congeners; thus, there are different relative proportions of congeners within given industrial mixtures. These mixtures exist as liquids to viscous solids. Between 1930 and 1977, when their industrial manufacture was banned in the United States, these mixtures were produced almost exclusively by Monsanto under the commercial name of Aroclors. Each Aroclor has a code number (e.g., Aroclor 1242, Aroclor 1248, and Aroclor 1254), the
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A Risk-Management Strategy for PCB-Contaminated Sediments FIGURE 2–2 Examples of PCB homologues.
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A Risk-Management Strategy for PCB-Contaminated Sediments FIGURE 2–3 Coplanar PCBs with no ortho-chlorines: (a) 3,4,4′,5-tetra-chlorinated biphenyl; (b) 3,3′,4,4′-tetra-chlorinated biphenyl; (c) 3,3′,4,4′,5,5′-hexa-chlorinated biphenyl; (d) 3,3′,4,4′,5-penta-chlorinated biphenyl; (e) comparison of the shape and size of a coplanar chlorinated biphenyl to 2,3,7,8-TCDD. last two numbers of which generally, but not always, refer to the percent by weight of chlorine in the mixture. For example, Aroclor 1254 is 54% chlorine by weight. Manufacturers of PCBs in other countries used different commercial names for PCBs—for example, Kanechlor (Japan), Santotherm (Japan), Phenolclor (France), Pyralene (France), Fenclor (Italy), Soval (Soviet Union), and Clophens (Germany). It is important to note that use of Aroclor as a trade name was not restricted to PCBs but was used for other polyhalogenated aromatic mixtures as well. History of PCBs PCBs, manufactured in the United States from 1929 to 1977, were used widely as insulating fluids in electrical equipment such as transformers and
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A Risk-Management Strategy for PCB-Contaminated Sediments capacitors, as well as in hydraulic systems, surface coatings, and flame retardants. Their chemical properties, such as nonflammability, chemical and thermal stability, dielectric properties, and miscibility with organic compounds, were responsible for many of their industrial applications. Their primary domestic uses in the United States as of 1970 are summarized in Table 2–1. Between 1929 and 1977, approximately 700,000 tons of PCBs were manufactured in the United States; 625,000 tons were used domestically and 75,000 tons were exported. Use of PCBs peaked in 1970 at 42,500 tons annually. The U.S. Environmental Protection Agency (EPA) estimates that over half of the PCBs sold in the United States were disposed of before enactment of federal regulations in 1976 (EPA 1999b). Although PCBs are no longer commercially manufactured and their disposal from existing industrial equipment is heavily regulated, there are several potential sources for continuing environmental releases. These sources include (1) continued use and disposal of PCB-containing products such as transformers, capacitors, and other electrical equipment that were manufactured before 1977, (2) combustion of PCB-containing materials, (3) recycling of PCB-contaminated products, such as carbonless copy paper, and (4) releases of PCBs from waste storage and disposal. Old consumer goods and household waste might also contain PCBs and their use and disposal are unregulated. The EPA database for registered electrical transformers (EPA 2000a) shows that, as of 1998, the 18,714 transformers listed contained a total of about 54,000 tons of PCBs, and as of 1988, 141,000 tons of PCBs remained in service in electrical equipment. Due to the long service life of this equipment, considerable amounts of PCBs are likely to remain in use for many TABLE 2–1 Domestic Uses of PCBs Category Type of Product New Total Use Closed electrical systems Transformer, capacitors, other (minor) electrical insulating and cooling applications 61% before 1971; 100% after 1971 Nominally closed systems Hydraulic fluids, heat transfer fluids, lubricants 13% before 1971; 0% after 1971 Open-end applications Plasticizers, surface coatings, ink and dye carriers, adhesives, pesticide extenders, carbonless copy paper, dyes 26% before 1971; 0% after 1971 Source: NRC (1979).
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A Risk-Management Strategy for PCB-Contaminated Sediments years. Spills of PCBs during handling or transport are an additional source of contamination. Between 1989 and 2001, there were 2,611 such spills (USCG 2001). Spills of greater than 1 pound of PCBs are reported to the EPA National Response Center. The National Toxics Inventory, an inventory conducted every 3 years by EPA under the Clean Air Act Amendments of 1990, reports atmospheric releases of hazardous air pollutants, including PCBs, from mobile and stationary sources. Point source air emissions of PCBs were 136 pounds per year from 127 maximum achievable control technology sources, which included utility boilers, industrial boilers, waste incineration, sewerage sludge incineration, portland cement manufacturing, municipal landfill, and other biological incineration (EPA 2000b). Unquantified emissions include accidental fires or uncontrolled combustion sources. The Toxic Release Inventory for 1998 reported that 3,747,166 pounds of PCBs were released (to air, surface water, land, and underground injection) from all industries in the United States (EPA 2000c). Facilities with effluent discharges are required to report PCB releases for permit compliance purposes under the Clean Water Act. On an annual basis, most of these releases are quite small. Under the Toxic Substances Control Act (TSCA) of 1976, all uses of PCBs are banned with certain exceptions. These exceptions include totally enclosed activities, such as certain electrical equipment—an authorized use, or exempted use under a special petition. These uses or activities are allowed because EPA has determined that they provide no unreasonable risk to human health or the environment. In general, materials containing PCBs at less than 50 parts per million (ppm) are considered “non-PCB” items by EPA and are not regulated under TSCA. Exemptions under TSCA for manufacturing, processing and distributing PCBs in commerce have been provided for their use in microscopy oils and research and development activities. TSCA also regulates the inadvertent generation of PCBs. EPA has estimated that more than 200 chemical processes can inadvertently generate PCBs. Products that might be a new source of PCBs include chlorinated solvents, paints, printing inks, agricultural chemicals, plastic materials, and detergent bars. The annual PCB concentration in wastes from these manufacturing processes or imported in the United States must average no more than 25 ppm, with a maximum level of 50 ppm; detergent bars must contain less than 5 ppm of PCBs. Releases of inadvertently produced PCBs from manufacturing operations must contain less than 10 ppm for air releases and less than 100 ppb for water discharges (EPA 1999b). Recently, a number of unrecognized uses of PCBs have been identified under the category of nonliquid PCBs (EPA 1999a). Currently in use, these solid materials were manufactured with PCBs as an intermediary reactant, and
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A Risk-Management Strategy for PCB-Contaminated Sediments include insulation (wool felt, foam rubber, and fiberglass), sound-dampening materials, paints, water-proofing materials, coatings for water pipes and storage tanks, and other materials, many of which have been found in federal buildings or military equipment, such as naval vessels. Although these solid materials might not present a current health risk from PCB exposure, they might become a significant source of PCB exposure as their utility expires. Another continuing source of PCBs are recycling activities that keep PCBs in circulation for many years. Materials that might contain PCBs include automobile and truck parts (e.g., nonmetallic parts such as glass and plastic), military equipment (e.g., ship parts), plastics, asphalt-roofing materials, and paper. In most other countries, PCB production is also banned. However, PCBs are reported to be manufactured in Russia and might also be manufactured in North Korea (Carpenter 1998). If that is the case, PCBs might be entering the environment both in those countries and in other countries that buy their PCB-containing products. Although these sources of PCBs are likely to be relatively small, they are a new source of PCBs in the environment. Unfortunately, estimates of continuing worldwide production of PCBs are not available. Such information would improve our understanding of the global balance of PCBs in the environment and the potential long-term impact of site management efforts. These continuing and new sources of PCBs to the global environment are important to consider as various physical, chemical, and biological processes transport PCBs regionally and globally. This issue is discussed in more detail in the next section. DISTRIBUTION AND DYNAMICS OF PCBs IN THE ENVIRONMENT The chemical properties of PCBs, such as stability and low reactivity, made them ideal for many industrial uses. PCBs are slow to biodegrade in the environment in comparison with many other organic chemicals and are generally persistent in all media. PCBs have relatively low water solubility and low vapor pressures (Erickson 1997) that allow them to partition between water and the atmosphere. Once released into the environment, PCBs tend to partition to the more organic components of the environment. For that reason, PCBs adsorb to organic matter in soils and sediments. As a result, PCBs can be found in almost every compartment in the environment (Tanabe 1988). PCBs adhere to the surfaces of organic particles in the water column, resulting in their eventual deposition and accumulation in sediments. The highest concentrations of PCBs are typically found in fine-grained, organically
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A Risk-Management Strategy for PCB-Contaminated Sediments rich sediments. Horizontal and vertical variations in PCB concentrations in sediments are common and are dependent on the history of PCB inputs to the ecosystem and on the temporal and spatial deposition patterns of fine- and coarse-grained sediments. At sites without new inputs of PCBs, the greatest concentrations tend to be found below the surficial sediments, where contaminated sediments are buried by less-contaminated sediments. The distribution of PCBs in sediments is affected by such factors as continuing use and disposal of PCBs; leaching from disposal sites; resuspension by turbulence; redeposition (hydrodynamic forces); chemical changes; and physical and biological mixing of the sediment. The different physical and chemical properties of the individual congeners determine their behavior during those various dynamic processes. As a result, identifying the specific environmental characteristics of PCB-contaminated sediments is challenging. Sediment characterization typically involves a combination of sampling techniques that include direct measurement of PCBs by high-resolution analytical methods and direct and indirect measurement of sediment properties. PCBs are considered to exist in three phases in the sediment and overlying water: freely dissolved, associated with dissolved organic carbon (DOC),1 and sorbed to particles.2 PCBs sorbed to particles are subject to settling, resuspension, and burial. Particles suspended in the water column are affected by hydrodynamic conditions. PCBs that are freely dissolved or associated with DOC can cross the sediment-water interface and move between the deeper sediments (below the bioturbation or bioactive surface sediment) and the surface sediment. This movement is largely a function of diffusion between the sediment pore water, and the overlying water column. It is dependent on the detailed hydrodynamic structure at the water-sediment interface and can be greatly enhanced by bioturbation caused by organisms living in the sediments. Freely dissolved PCBs in the water column are also subject to volatilization across the air-water interface. Such loss can be substantial, especially in systems that provide substantial time for the water-air interactions. Transformations of PCBs can also occur in aquatic systems by microbial degradation (in aerobic water columns and surficial sediments), reductive dechlorination (in anaerobic sediments), and metabolism via organisms that 1 The term, “associated with DOC” is used, because the exact mechanism of interaction of PCBs with DOC is not well-defined. DOC can include colloidal materials that are mostly organic matter. 2 The term “sorbed” suggests that a combination of adsorptive and absorptive processes are involved, depending upon the types of particles.
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A Risk-Management Strategy for PCB-Contaminated Sediments take up the PCBs. Metabolism by microorganisms (Mavoungou et al. 1991) and animals (McFarland and Clarke 1989) can cause relative proportions of some congeners to increase while others decrease (Boon and Eijgenraam 1988; Borlakoglu and Walker 1989). Because the susceptibility of PCBs to degradation and bioaccumulation is congener specific, the composition of PCB congener mixtures that occur in the environment differs substantially from that of the original industrial mixtures released into the environment (Zell and Ballschmiter 1980; Giesy and Kannan 1998; Newman et al. 1998). In addition to environmental transformation products of PCBs, other chemicals, such as polyaromatic hydrocarbons (PAHs), polychlorinated dibenzofurans (PCDFs), polychlorinated dibenzo-p-dioxins (PCDDs), pesticides, and metals, might be present in contaminated sediments. Generally, the less-chlorinated congeners are more water soluble, more volatile, and more likely to biodegrade. Therefore, lower concentrations of these congeners are found in sediments compared with the original concentrations of Aroclors that entered the environment. Higher-chlorinated PCBs are often more resistant to degradation and volatilization and sorb more strongly to particulate matter. Some of these more-chlorinated PCBs tend to bioaccumulate to greater concentrations in tissues of animals than do lower-molecular-weight PCBs. The more-chlorinated PCBs can also biomagnify in food webs (see Box 2-1), and other higher-molecular-weight congeners have specific structures that make them susceptible to metabolism by enzymes once these congeners are taken up by such species as fish, crustacea, birds, and mammals. The low vapor pressure of PCBs, coupled with air, water, and sediment transport processes, means that they are readily transported from local or regional sites of contamination to remote areas (Risebrough et al. 1968; NRC 1979; Atlas and Giam 1981; Subramanian et al. 1983). PCBs can enter a global biogeochemical cycle that transports them far from their initial source of input. This global biogeochemical cycling of PCBs is the result of volatilization losses from tropical and subtropical waters to the atmosphere. These atmospheric PCBs move from warmer regions to polar regions, especially in the northern hemisphere, where they are deposited to soil and water surfaces (Muir et al. 2000). Table 2–2 presents some atmospheric concentrations of PCBs from various regions of the world, illustrating the scale and variability of their global distribution. POTENTIAL EXPOSURE PATHWAYS Humans and wildlife can be exposed to PCBs either directly from contact with contaminated air, sediments, or water or indirectly through the diet.
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A Risk-Management Strategy for PCB-Contaminated Sediments BOX 2–1 Definitions Bioaccumulation—The net accumulation of PCBs by an organism as a result of uptake from all routes of exposure (i.e., water, sediment, food, or air) (Suter 1993). Bioconcentration—The net accumulation of PCBs directly from water by aquatic organisms (Suter 1993). Food Web Transfer—The movement of PCBs in the tissue of prey to the tissue of the predator, repeated one or more times in the food web, where the predator of the first transfer is the prey in the next step (Van Leeuwen and Hermens 1995). Biomagnification—The tendency of PCBs to accumulate to higher concentrations at higher levels in the food web through dietary accumulation (Suter 1993). Bioavailability—The ratio of the amount of PCBs taken into the organism and thus available to internal tissues, compared to the amount of PCBs ingested into the gut, inhaled into the lungs, or in direct contact with the skin (Suter 1993). When considering exposure pathways, it is imperative to assess the biologically available fraction of PCBs. In sediments, PCBs can be buried below the biologically active zone and, therefore, are less available for uptake by aquatic organisms. The biologically active zone is the top layer of sediments, typically 5–10 centimeters (cm) deep. This layer is continuously reworked by sediment-dwelling organisms and remains in contact with the overlying water. PCBs that are strongly sorbed to organic sediment particles in the biologically active zone tend to have reduced bioavailability to organisms that ingest or are exposed to these sediments (EPA 1994). Consumption of PCB-contaminated foods is the most significant route of exposure to PCBs for the general human population (Newhook 1988; Birmingham et al. 1989; Fitzgerald et al. 1996). This exposure occurs as a result of bioaccumulation of PCBs through the food chain. For example, PCBs can enter the aquatic food web via uptake by benthic invertebrates that are in close contact with the contaminated sediments. These invertebrates are eaten by other aquatic organisms, such as fish, and thus the PCBs migrate up the food
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A Risk-Management Strategy for PCB-Contaminated Sediments TABLE 2–2 Global Atmospheric PCB Concentrations in Ambient Outdoor Air Location Concentration (ng/m3) Reference Antarctic coast 0.06–0.2 Tanabe et al. 1983 Canadian Arctic (81°N) 0.1–0.3 Bidleman et al. 1990 Remote 0.02–0.5 Eisenreich et al. 1983 Great Lakes 0.1–5 Eisenreich et al. 1983 Rural 0.1–2 Eisenreich et al. 1983 Urban 0.5–30 Eisenreich et al. 1983 Lake Superior, U.S. (peak in spring) 0.2 Hornbuckle et al. 1994 Lake Superior, U.S. (low point in fall) 0.065 Hornbuckle et al. 1994 Various U.S. locations 0.02–36 NRC 1979 Marine air 0.05–2.0 NRC 1979 Atlantic Ocean 0.05 WHO 1976 Gulf of Mexico 0.2–0.9 Giam et al. 1976 North Pacific Ocean 0.54 Atlas and Giam 1981 North Atlantic Ocean 1.84 Tanabe at al. 1982 West Pacific Ocean 0.06–1.2 Tanabe et al. 1982 Bermuda 0.4 Panshin and Hites 1994a Bloomington, IN 0.7–2.5 Panshin and Hites 1994b North Atlantic Ocean <0.05–1.6 Bidleman et al. 1976, 1981 Lake Baikal, Siberia 0.009–0.023 Iwata et al. 1995 Several oceans and seas 0.004–0.6 Iwata et al. 1993 Arctic 0.002–0.013 (other studies also reviewed) Bidleman et al. 1990 Tokyo, Japan 20 Kimbrough 1980 Matsuyama, Japan 2–5 Kimbrough 1980 Sweden <0.8–3.9 WHO 1976 Germany 5–10 Benthe et al. 1992 United States 5 WHO 1976 Landfills, U.S. 2–18 MacLeod 1979 Electrical substations, U.S. 1–47 MacLeod 1979 Transformer manufacturer, U.S. 17–5,900 MacLeod 1979 Spill site, U.S. 10–10,800 MacLeod 1979 Source: Adapted from Erickson (1997).
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A Risk-Management Strategy for PCB-Contaminated Sediments species. Many of the declines might be caused by eating contaminated fish. Although declines were initially likely to be caused by DDE-induced toxicity, the populations still exhibit subtle effects, such as deformities that might be caused by dioxin-like PCBs (Giesy et al. 1994,a,b). It has been proposed that PCB concentrations in tree swallows nesting along the Hudson River might be responsible for their reduced reproductive success (Secord and McCarty 1997; Secord et al. 1999). In some locations, for example, the Niagara River, when wildlife health criteria values have been calculated and compared with the concentrations of PCBs found in local fish species, the amounts of PCBs in the fish have exceeded the criteria values (Newell et al. 1987). The ecological impacts of PCBs on wildlife have also been assessed in the laboratory. Most of the laboratory studies on wildlife have confirmed the field work. However, these studies are problematic as controlled laboratory conditions cannot be used directly to predict effects in real-world populations because of changes in the concentrations and composition of PCBs as a function of space and time. Thus, the PCB mixture to which organisms are exposed at one time or at one location might be very different from that to which they are exposed in the laboratory or at other times or locations in the field. The pattern of relative proportions of PCBs in environmental mixtures is variable and does not resemble the composition of the original technical PCB mixtures that were released into the environment (Kannan et al. 1993; Corsolini et al. 1995a,b). Furthermore, the relative concentrations of various PCB congeners differ according to trophic level and species. WEATHERING OF PCBs The compositions of PCB congener mixtures that occur in the environment differ substantially from those of the original technical Aroclor mixtures released to the environment (Zell and Ballschmiter 1980; Giesy and Kannan 1998; Newman et al. 1998). As discussed previously, the difference is due to the changes in the composition of PCB mixtures over time after release into the environment because of several processes collectively referred to as “environmental weathering.” The weathered multicomponent mixtures might have significant differences compared with Aroclor standards; the degree and position of chlorine substitution not only influences the physical and chemical properties of the PCB congeners but also their toxic effects. Weathering is a result of the combined effects of such processes as differential volatilization, solubility, sorption, anaerobic dechlorination, and metabolism, and results in changes in the composition of the PCB mixture over time and between trophic levels (Froese et al. 1998). Less-chlorinated PCBs are often lost rapidly due
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A Risk-Management Strategy for PCB-Contaminated Sediments to volatilization and metabolism, whereas more-chlorinated PCBs are often resistant to degradation and volatilization and sorb more strongly to particulate matter. Bioaccumulation in the tissues of animals is greater for more-chlorinated PCBs than for less-chlorinated PCBs; therefore, more-chlorinated PCBs are more likely to biomagnify in food webs. Microbial reductive dechlorination of PCBs is a process shown to occur in a variety of anaerobic environments (Bedard and Quensen 1995). This process does not remove all the chlorines and does not alter the basic structure of the biphenyl. The process results in a decrease in the concentrations of some congeners and an increase in others; therefore, the change in the total molar concentration of PCBs in sediments is generally not great. Reductive dechlorination occurs preferentially for chlorines in the meta and para positions, thereby selectively reducing the relative proportions of the PCBs that are laterally substituted. These congeners are also those that tend to have the greatest potency to cause AhR-mediated effects. One of the most potent of the laterally substituted, non-ortho-substituted congeners is congener 126 (3,3′,4,4′,5-pentachlorobiphenyl). The absolute and relative concentration of this congener was reported to decrease by as much as 10- to 100-fold due to reductive dechlorination (Quensen et al. 1998). The total concentration of TCDD-toxicity equivalents (TEQ) in sediments (see Chapter 6 for a discussion of TEQ and toxicity equivalence factors (TEF)), either determined by the H4IIE bioassay or by application of TEFs to concentrations of individual congeners, was reduced 100-fold by reductive dechlorination (Quensen et al. 1998). Because the TEQ of the total PCB mixture has been shown to be the critical toxicant and the most predictive of toxicity of environmental mixtures of PCBs (Giesy and Kannan 1998; van den Berg et al. 1998), the TEQ reduction suggests that the toxicity of the total PCB mixture would be reduced by approximately 100-fold (Quensen et al. 1998). Furthermore, the bioavailability of the AhR-active congeners has been shown to be less than that for the di-ortho-substituted congeners. Thus, both processes, reductive dechlorination and selective sorption of coplanar PCB congeners, tend to reduce the toxicity of the mixture, relative to technical Aroclor mixtures, during the weathering process. As was discussed above, the most accurate method of estimating the relative toxic potency of PCB mixtures is to measure the concentrations of individual congeners in tissues of receptors and correct their toxic potency by use of toxic potency factors. It is not appropriate to use thermodynamic models to predict the movement of total PCB or TEQ concentrations from one matrix or trophic level to another. The movement of individual congeners, or at least those with more similar partitioning characteristics, should be modeled and the congeners should be corrected for their toxic potency. Thus, to model
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A Risk-Management Strategy for PCB-Contaminated Sediments the toxicity of the complex PCB mixture that is in sediments to higher trophic levels requires the application of both TEFs and congener-specific biomagnification factors (BMFs). When the toxicity of an example set of congener-specific concentrations in fish tissues to mink was estimated, it was found that the critical toxicant was the TEQ (Foley et al. 1988). CONCLUSIONS AND RECOMMENDATIONS PCBs are complex mixtures of chemicals that can have adverse effects on humans and wildlife. The committee’s review of recent scientific information supports the conclusion that exposure to PCBs might result in chronic effects—such as cancer and immunological, developmental, reproductive, neurological effects—in wildlife, laboratory animals, and possibly humans. Therefore, the committee considers the presence of PCBs in sediments to pose potential long-term public health and ecosystem risks. It must be understood by all affected parties that even if the risks at a site are managed such that a specific sediment concentration of PCBs is achieved, over time PCB concentrations will slowly change due to numerous factors including atmospheric inputs from other sources and biodegradation. Although considerable new information has become available in the past 2 decades and new information from field and laboratory studies is reported regularly, the committee finds that further research is particularly warranted in the following areas: Additional data are needed on the toxicological effects of exposure to multiple chemicals—PCBs plus PAHs, PCDDs, and metals—and to “real-world” mixtures of PCBs. To collect such data, further elucidation of the various mechanisms of toxic actions will be required. A better understanding of the contribution of PCB-contaminated sediments to the total global burden of PCBs is needed. The role of global cycling of PCBs in assessing the PCB problem at a specific site should be considered. REFERENCES Amaro, A.R., G.G.Oakley, U.Bauer, H.P.Spielmann, and L.W.Robertson. 1996. Metabolic activation of PCBs to quinones: reactivity toward nitrogen and sulfur nucleophiles and influence of superoxide dismutase. Chem. Res. Toxicol. 9(3):623–629.
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Representative terms from entire chapter: