Non-Dioxin-Like Polychlorinated Biphenyls
Polychlorinated biphenyls (PCBs) are a class of chlorinated aromatic hydrocarbon chemicals that once were used as heat-exchanger, transformer, and hydraulic fluids, and as additives to paints, oils, joint caulking, and floor tiles. Peak production occurred in the early 1970s, and production was banned in the United States after 1979. More than 1.5 billion pounds of PCBs were manufactured in the United States prior to 1977. The continued concern about these chemicals is because of their persistence in the environment and accumulation in wildlife and the animal food chain.
Food is the main source of exposure for the general population. PCBs enter the food chain by a variety of routes, including migration into food from external sources, contamination of animal feeds, and accumulation in the fatty tissues of animals. PCBs are found at higher concentrations in fatty foods (e.g., dairy products and fish). The transfer of PCBs from mother to infant via breast milk is another important source of exposure. The lesser-chlorinated PCBs are more volatile and indoor inhalational exposure from buildings containing caulking made with these PCBs prior to 1979 can increase background serum levels (Johansson et al., 2003; Kohler et al., 2005). Other sources of exposure in the general population include the release of these chemicals from PCB-containing waste sites and from fires involving transformers and capacitors. Additionally, the heat from fires can result in the production of polychlorinated dibenzofurans from PCBs. In certain occupational settings, workers can be exposed to PCBs when repairing or manufacturing transformers, capacitors, and hydraulic systems, and when remediating hazardous-waste sites. Both U.S. FDA and OSHA have developed criteria on the allowable levels of these chemicals in foods and the workplace. The U.S. EPA has also set criteria for allowable levels in water and waste materials. The international Stockholm Convention on Persistent Organic Pollutants of 2001 establishes the most stringent guidelines to date regarding elimination, restriction and unintentional production of PCBs and selected organochlorine chemicals (Porta and Zumeta, 2002).
Exposure to these chemicals nearly always occurs as mixtures rather than from individual PCBs. The different types of PCB chemicals are known as congeners, which are compounds that are distinguished by the number of chlorine atoms and their location on the biphenyl structure. PCB congeners can be divided into the coplanar, the mono-ortho-substituted PCBs, and other non-dioxin-like PCBs. The significance of this designation is that the coplanar and some of the mono-ortho-substituted PCBs have dioxin-like toxicologic effects. Structural nomenclature is available at: http://www.epa.gov/oswer/riskassessment/pdf/1340-erasc-003.pdf. The non-dioxin-like PCBs and their metabolites do not interact substantially with the aryl hydrocarbon receptor (AhR) and may act through different pathways than the dioxin-like chemicals, so their effects are not represented in the use of toxic equivalency factors (TEFs) (Carpenter, 2006). Commonly measured non-dioxin-like PCBs are listed in the table.
|Non-Dioxin-like Polychlorinated Biphenyls Measured in the National Biomonitoring Program|
|Non-dioxin-like polychlorinated biphenyls (IUPAC number)||CAS number|
|Polychlorinated biphenyls (general class)||1336-36-3|
|2,4,4'-Trichlorobiphenyl (PCB 28)||7012-37-5|
|2,2',3,5'-Tetrachlorobiphenyl (PCB 44)||41464-39-5|
|2,2',4,5'-Tetrachlorobiphenyl (PCB 49)||41464-40-8|
|2,2',5,5'-Tetrachlorobiphenyl (PCB 52)||35693-99-3|
|2,3',4,4'-Tetrachlorobiphenyl (PCB 66)||32598-10-0|
|2,4,4',5-Tetrachlorobiphenyl (PCB 74)||32690-93-0|
|2,2',3,4,5'-Pentachlorobiphenyl (PCB 87)||38380-02-8|
|2,2',4,4',5-Pentachlorobiphenyl (PCB 99)||38380-01-7|
|2,2',4,5,5'-Pentachlorobiphenyl (PCB 101)||37680-73-2|
|2,3,3',4',6-Pentachlorobiphenyl (PCB 110)||38380-03-9|
|2,2',3,3',4,4'-Hexachlorobiphenyl (PCB 128)||38380-07-3|
|2,2',3,4,4',5'-Hexachlorobiphenyl (PCB 138)||35065-28-2|
|2,3,3',4,4',6-Hexachlorobiphenyl (PCB 158)||74472-42-7|
|2,2',3,4',5,5'-Hexachlorobiphenyl (PCB 146)||51908-16-8|
|2,2',3,4',5',6-Hexachlorobiphenyl (PCB 149)||38380-04-0|
|2,2',3,5,5',6-Hexachlorobiphenyl (PCB 151)||52663-63-5|
|2,2',4,4',5,5'-Hexachlorobiphenyl (PCB 153)||35065-27-1|
|2,2',3,3',4,4',5-Heptachlorobiphenyl (PCB 170)||35065-30-6|
|2,2',3,3',4,5,5'-Heptachlorobiphenyl (PCB 172)||52663-74-8|
|2,2',3,3',4,5',6'-Heptachlorobiphenyl (PCB 177)||52663-70-4|
|2,2',3,3',5,5',6-Heptachlorobiphenyl (PCB 178)||52663-67-9|
|2,2',3,4,4',5,5'-Heptachlorobiphenyl (PCB 180)||35065-29-3|
|2,2',3,4,4',5',6-Heptachlorobiphenyl (PCB 183)||52663-69-1|
|2,2',3,4',5,5',6-Heptachlorobiphenyl (PCB 187)||52663-68-0|
|2,2',3,3',4,4',5,5'-Octachlorobiphenyl (PCB 194)||35694-08-7|
|2,2',3,3',4,4',5,6-Octachlorobiphenyl (PCB 195)||52663-78-2|
|2,2',3,3',4,4',5,6'-Octachlorobiphenyl (PCB 196)||42740-50-1|
|2,2',3,3',4,5,5',6'-Octachlorobiphenyl (PCB 199)||52663-75-9|
|2,2',3,4,4',5,5',6-Octachlorobiphenyl (PCB 203)||52663-76-0|
|2,2',3,3',4,4',5,5',6-Nonachlorobiphenyl (PCB 206)||40186-72-9|
|2,2'3,3'4,4'5,5'6,6'-Decachlorobiphenyl (PCB 209)||2051-24-3|
Human health effects that have been reported after investigations of occupational and accidental exposures to high levels of PCBs include elevations of serum hepatic enzymes, dermal changes, inconsistent associations with serum lipid levels, and some types of cancer (e.g., liver, biliary) (ATSDR, 2000; Carpenter, 2006; Charles et al., 2001; Negri et al., 2003). Animal studies have demonstrated varied effects of PCBs including neurotoxicity, immune suppression, altered thyroid and reproductive function, and liver cancer (Carpenter, 2006; U.S.EPA, 2008). Effects of PCBs in humans are difficult to study because of co-exposures to the dioxin-like chemicals and other organochlorine chemicals.
Transplacental transfer of PCBs after maternal environmental exposure has been reported to be associated with altered psychomotor development in children and lower birth weight and size in newborns ( Hertz-Picciotto et al., 2005; Jacobson and Jacobson, 1996; Koopman-Essenboom et al., 1996; Longnecker et al., 2003; Lundqvist et al., 2006; Sagiv et al., 2007; Sala et al., 2001), although other studies have either not confirmed these findings or found that such effects do not persist into toddler and school aged children (Gladen and Rogan, 1991; Gray et al., 2005; Hertz-Picciotto et al., 2005; Koopman-Essenboom et al., 1996; Wolff et al., 2007). Many animal studies demonstrate that high dose PCB impairs neurodevelopment or their hydroxylated metabolites may interfere with thyroid hormone-dependent neurodevelopment (Kimura-Kuroda et al., 2007; Nguon et al., 2005; Purkey et al., 2004; Roegge et al., 2006).
The non-dioxin-like PCBs weakly interact with estrogen and thyroid receptors and with transport proteins, and the hydroxylated metabolites of PCBs may be more potent mediators of these actions (Azulmozhiraja et al., 2005; DeCastro et al., 2006; Langer et al., 2005; Purkey et al., 2004; Kitamura etal., 2005; You et al., 2006). Variations in thyroid hormone levels have been associated with PCB exposures in human populations (Langer et al., 2007a; Meeker et al., 2007; Otake et al., 2007; Wang et al., 2005). Though only limited investigation of estrogenic or reproductive effects has occurred in women, inconsistent associations of PCB levels with altered spermatogenesis and reproductive hormone levels have been reported in environmentally exposed men (Giwercman et al., 2006; Rignell-Hydbom et al., 2005; Toft et al., 2006).
PCBs are not considered directly genotoxic. They are classified as probable human carcinogens by IARC and are classified by NTP as reasonably anticipated to be carcinogens. Early studies associated workplace PCB exposures with increased deaths from cancer of the liver, gallbladder, biliary tract, gastrointestinal tract, brain and malignant melanoma (Knerr and Schrenk, 2006). Follow up studies of these earlier investigations have shown no increase in deaths or cancers, with the exception of liver cancer (Kimbrough et al., 2003; Prince et al., 2006; Ross, 2004), though the contributions of dioxin-like chemicals or other organochlorines were unclear. Recent studies have associated PCB exposures with other cancers (De Roos et al., 2005; Engel et al., 2007; Prince et al., 2006). Information about external exposure (i.e., environmental levels) and health effects is available from ATSDR at http://www.atsdr.cdc.gov/toxprofiles/index.asp and from the U.S. EPA at http://www.epa.gov/iris.
Measurement of serum PCBs generally reflect cumulative past exposure. Levels of non-dioxin-like PCBs in NHANES 2003-2004 are observed to be roughly similar to the previous two NHANES survey periods CDC, 2013). Many PCBs can remain in the body for years after exposure, though some of the PCBs with fewer chlorine atoms have short residence times. The levels of individual PCB congeners in the body may vary by exposure source and by differences in pharmacokinetics, i.e., those with longer half-lives accumulate to higher levels. Adult age-related accumulations in the non-dioxin-like PCBs have been observed in many studies (Apostoli et al., 2005; Park et al., 2007; Patterson et al., 1994). Breastfeeding is a major source of PCBs, with serum levels increasing after birth in breastfed infants and then decreasing in early adolescence due to dilution as body mass increases (Barr et al., 2006). Fish consumption from the Great Lakes region contributed a twofold to tenfold increase in the mean concentrations of non-dioxin-like PCBs over referent populations (Patterson et al., 1994; Turyk et al., 2006). Arctic native Alaskans who consumed locally-caught fish, meat, and eggs had mean serum levels of total PCB that were nearly three times higher than the adult NHANES 1999-2000 subsample (Carpenter et al., 2005; CDC, 2013; Needham et al., 2005). Much higher levels due to contaminated fish intake have also been noted in eastern Europe (Langer et al., 2007b).
The concentrations of the di-ortho-substituted PCBs were usually higher than the mono-ortho-substituted PCBs, which in turn were higher than the coplanar PCBs (CDC, 2013; Glynn et al., 2000; Longnecker et al., 2000; Patterson et al., 1994). The most frequently detected di-ortho-chlorine-substituted PCBs in population studies are 138, 153, and 180 (CDC, 2013; Heudorf et al., 2002; Patterson et al., 1994, 2009; Turyk et al., 2006). These three congeners contributed a substantial portion of the total PCB concentration observed in pooled specimens representative of a New Zealand population (Bates et al., 2004); in a small population of Swedish men (Glynn et al., 2000), and in blood bank specimens from Canada (Longnecker et al., 2000). In the U.S representative subsample from NHANES 1999-2000, non-dioxin-like PCBs 138, 153, and 180 accounted for 65% of the measured total sum of PCBs (Needham et al., 2005) and for 78% of the total in a referent population of 311 Italian residents in 2001-2003 (Apostoli et al., 2005). Non-dioxin-like PCBs with five, six, and seven chlorines attached comprised about 80% of the total PCBs in human serum, or alternatively, PCBs 138, 153, 180, 187 and 118 composed 57% of the total PCB concentration in a small sample of South Korean residents and incineration workers (Park et al., 2007). In the sera of Yucheng victims analyzed 15 years after the rice oil contamination event, 73% of the total PCB concentration was contributed by PCBs 99, 138, 153, 156, 170, 179, and 180 (Hsu et al., 2005).
Median serum lipid-adjusted levels of PCB 153 declined by 38% from 1991 to 2001 in a small sample of Swedish men (Hagmar et al., 2006). In four biannual surveys covering the years 1996-2003, about 400 German fourth grade children were sampled each period and demonstrated a decrease of more than one-half in mean whole blood levels of PCBs 138, 153, and 180 (Link et al., 2005). Lipid adjusted levels of the non-dioxin-like PCBs seen in the U.S representative subsamples from NHANES 2001-2002 were generally lower than levels in selected populations during the 1980s to 1990s (CDC, 2013; Glynn et al., 2000; Longnecker et al., 2000; Patterson et al., 1994).
In a convenience sample of 624 urban Germans aged 0-65 years conducted during 1998 (Heudorf et al., 2002), 95th percentile levels for PCBs 138, 153, and 180 were similar or up to two times higher than 95th percentile levels in the NHANES 1999-2000 subsample (CDC, 2013). In contrast, levels in pooled serum samples from a representative population of New Zealand residents in 1996-1997 were slightly lower than NHANES 1999-2000 (Bates et al., 2004). In two separate Italian studies of a regional reference population and a convenience sample in 2001-2003, median serum levels of PCBs 138, 153, and 180, as well as the sum of measurable PCBs, were about five times higher than NHANES 1999-2000 (Apostoli et al., 2005; CDC, 2013; Needham et al., 2005; Turci et al., 2006). The mean levels of PCBs 153 and 180 in 753 adult native Americans were approximately similar to the 95th percentile for the overall adult NHANES 2001-2002 population (CDC, 2013; DeCaprio et al., 2005). In some other countries, comparable population levels were ten or more times higher than those reported for NHANES subsamples from 1999-2000 and 2001-2002 (CDC, 2013; Jursa et al.,2006; Petrik et al., 2006). In the sera of Yucheng victims analyzed 15 years after the contamination event, mean serum lipid adjusted levels of PCBs 99, 153, 170, and 180 were up to eight times higher than the 95th percentiles of NHANES 1999-2000 (Hsu et al., 2005).
Finding a measurable amount of one or more PCBs in serum does not mean that the levels of the PCBs cause an adverse health effect. Biomonitoring studies of serum PCBs can provide physicians and public health officials with reference values so that they can determine whether or not people have been exposed to higher levels of PCBs than levels found in the general population. Biomonitoring data can also help scientists plan and conduct research on exposure and health effects.
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