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«Ruder AM. Potential health effects of occupational chlorinated solvent exposure. Ann NY Acad Sci 2006; 1076: 207-227. Schoeny R, Haber L, Dourson M. ...»

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Ruder AM. Potential health effects of occupational chlorinated solvent exposure. Ann NY

Acad Sci 2006; 1076: 207-227.

Schoeny R, Haber L, Dourson M. Data considerations for regulation of water contaminants.

Toxicology 2006; 221: 217-224.

Tan YM, Liao KH, Clewell 3rd HJ. Reverse dosimetry: interpreting trihalomethanes

biomonitoring data using physiologically based pharmacokinetic modeling. J Expo Sci

Environ Epidemiol 2007; 17: 591-603.

Tan YM, Liao KH, Conolly RB, et al. Use of a physiologically based pharmacokinetic model to identify exposures consistent with human biomonitoring data for chloroform. J Toxicol Environ Health A 2006; 69: 1727-1756.

Villanueva CM, Cantor KP, Cordier S, et al. Disinfection byproducts and bladder cancer.

Epidemiology 2004; 15: 357-367.

Villanueva CM, Cantor KP, Grimalt JO, et al. Bladder cancer and exposure to water disinfection by-products through ingestion, bathing, showering, and swimming in pools. Am J Epidemiol 2007; 165: 148-156.

Wang GS, Deng YC, Lin TF. Cancer risk assessment from trihalomethanes in drinking water.

Sci Total Environ 2007; 387: 86-95.

Weinberg HS, Kranser SW, Richardson SD, Thurston Jr AD. The occurrence of disinfection by-products (DBPs) of health concern in drinking water: results of a nationwide DBP occurrence study. 2002. EPA/600/R02/068.

Xu X, Weisel CP. Dermal uptake of chloroform and haloketones during bathing. J Expo Anal Environ Epidemiol 2005; 15: 289-296.

Yamamoto S, Kasai T, Matsumoto M, et al. Carcinogenicity and chronic toxicity in rats and mice exposed to chloroform by inhalation. J Occup Health 2002; 44: 283-293.

Zwiener C, Richardson SD, DeMarini DM, et al. Drowning in disinfection byproducts?

Assessing swimming pool water. Environ Sci Technol 2007; 41: 363-372.

Polychlorinated biphenyls (PCBs) by Larry W. Robertson PhD and Avima Ruder PhD Citation for most recent IARC review IARC Supplement 7, 1987 In IARC Monograph Volume 100 F (in press), PCB 126 is carcinogenic to humans (Group 1) (Baan et al 2009). In making the evaluations, the Working Group considered the following mechanistic arguments: There is strong evidence to support a receptor-mediated mechanism for PCB 126 carcinogenesis in humans based upon evidence of carcinogenicity in experimental animals and upon extensive evidence showing activity identical to 2,3,7,8TCDD for every step of the mechanism described for 2,3,7,8-TCDD carcinogenesis in humans including receptor binding, gene expression, protein activity changes, cellular replication, oxidative stress, promotion in initiation-promotion studies and complete carcinogenesis in laboratory animals. (http://monographs.iarc.fr/ENG/Meetings/vol100Fevaluations.pdf) Current evaluation

Conclusion from the previous review:

There is limited evidence for the carcinogenicity to humans of polychlorinated biphenyls (Group 2A). The available studies suggest an association between cancer and exposure to PCBs. The increased risk from hepatobiliary cancer emerged consistently in different studies.

Since, however, the numbers were small, dose-response relationships could not be evaluated, and the role of compounds other than PCBs could not be excluded, the evidence was considered to be limited.

Environmental Exposure and Biomonitoring PCBs were widely used from the 1930’s through the 1980’s and later, with an estimated total production of about 1.3 million metric tons (Breivik et al., 2002). Exposure continues from leaks from transformers and capacitors, volatilization of PCBs in cites, in buildings, from sewage, landfills and waste sites, and combustion of materials containing PCBs (Dyke et al., 2003).

Occupational Exposure PCB exposures associated with occupational settings have greatly diminished since the 1970’s, due to the ban on new uses for PCBs. Since the production of PCBs ended worldwide in 1993 (Breivik et al., 2002), new occupational exposure has been confined to four groups of workers: personnel replacing or repairing transformers and capacitors still containing PCB dielectric fluid (Altenkirch et al., 1996; Hay and Tarrel 1997; Shalat et al., 1989; Wolff et al., 1992); first responders to incidents where a transformer has exploded (Kelly et al., 2002);

construction workers removing old paint, plaster, caulk, or floor finishes containing PCBs (Fromme et al., 1996; Herrick et al., 2004; Herrick et al., 2007; Piloty and Koppl 1993; Rudel and Perovich 2009; Rudel et al., 2008), and workers at hazardous waste disposal sites (Gonzalez et al., 2000; Zhao et al., 2006). The serum levels of workers engaged in sealant removal were 2-10 times higher at the end of these activities than they had been one year before (Kontsas et al., 2004). Workers exposed occupationally while PCBs were still in general use have a body burden of PCBs from their former exposure (Schecter et al., 2002;

Wolff 1985). A large number of capacitors and transformers filled with PCBs is still in use (Environmental Protection Agency and Environment Canada 2005) so potential occupational exposure continues.

Foodborne Exposure Food chain exposure incidents by accident or malice have also occurred; one is reminded of Yusho (1968), Yu-Cheng (1979), the French cheese contamination episode (1976) and the Belgium “Dioxin” Crisis of 1999 (Covaci et al., 2008). Lower levels of PCBs are broadly prevalent in foodstuffs, and a great deal of attention has been paid to these PCBs, especially higher chlorinated ones, those more resistant to metabolic transformation, and those from certain food sources, especially in fish (Ludewig et al., 2007). People living on or near PCBcontaminated soil or near PCB-contaminated water, those eating contaminated foods, and those living in old homes being renovated continue to be exposed (Patterson et al., 2009;

Rudel et al., 2008; Weintraub and Birnbaum 2008; Zheng et al., 2008).

Airborne Exposure Air as a source of environmental PCB exposure was nearly completely ignored until about a decade ago. Systematic measurements of atmospheric PCBs started only in the 1990’s. The first urban monitoring site was installed in Chicago in 1995. The level of PCB contamination in the air is strongly influenced by temperature. In Chicago air concentrations between 100-300 pg/m3 in winter and up to 5,000-16,000 pg/m3 on hot summer days were reported (Green et al., 2000).

Inhalation exposure is considered to be a major route of occupational exposure to PCBs, and it was estimated that in capacitor workers, for example, a maximum of 80% of adipose PCBs may have been absorbed by inhalation exposure (Wolff 1985). Recently even higher levels of PCBs were measured in indoor air in buildings constructed in the 1970’s using joint sealants that contained 4-9% PCBs. Indoor air concentrations up to 13,000 ng/m3 were measured in some classrooms of a contaminated school (Kohler et al., 2002), which is more than an order of magnitude above the NIOSH guideline of 1 μg/m3 (NIOSH, 2004) for occupational settings.

Other possible sources for indoor PCBs are believed to be data screen terminals (Digernes and Astrup 1982), ceiling tiles and fluorescent light capacitors (Harris 1985). It was reported that the concentration of PCBs in indoor air can be at least an order of magnitude higher than in outdoor air (Balfanz et al., 1993; Vorhees et al., 1997; Wallace et al., 1996); however, regional outdoor levels can be very high due to activities like building renovations, dredging, or contamination from cement factory exhaust. Thus under certain circumstances the intake from inhalation exposure exceeds PCB intake from food.

PCBs in foods, like fish or mothers’ milk, and in human adipose tissue are usually the higher chlorinated ones, where congeners like PCB153, PCB180, PCB183 and others predominate.

Airborne PCBs are very different, since they require volatilization. Major congeners in Chicago air are PCB5/PCB8 (co-elute), PCB18, PCB28, PCB44, PCB52, PCB77/PCB110 (co-elute), PCB95, PCB101, to name a few (Hu et al., 2008; Zhao et al., 2009). Of two populations in Italy the more urban group had significantly higher levels of lower chlorinated PCBs (PCB52 was about 100-fold higher) than the population in a more rural environment (Turci et al., 2006). In Germany, PCB28 and PCB52 were the prevailing congeners in indoor air of contaminated schools (Kohler et al., 2002; Schwenk et al., 2002). Elevated levels of PCB28 and PCB52 were measured in the blood of teachers from these schools compared to non-contaminated schools, whereas the mean blood levels of higher chlorinated PCBs, i.e. PCB138, PCB153 and PCB180 were almost identical (Schwenk et al., 2002). Children in schools with 690-20,800 ng PCB/m3 air had median levels of 6, 9, and 5 ng/l PCB28, PCB52, and PCB101, respectively, whereas children in non-contaminated schools had levels below the detection level of 1 ng/L (Liebl et al., 2004). Both groups had no significant differences in PCB138, PCB153, and PCB180 levels, indicating that indoor air exposure contributed to the PCB body burden. In Germany the nonoccupational tolerable indoor total PCB concentration was limited to 300 ng/m3 (PCB Guideline

1995) based on a tolerable daily intake (TDI) of a total of 1 µg/kg body weight. Not only were these levels exceeded in several schools, but this TDI was based on a chronic toxicity study with a commercial PCB mixture, which measured hepatic enzyme induction as endpoint (Chen and Dubois 1973). Airborne PCB profiles are distinctly different from those of commercial PCB mixtures like Aroclor 1254, and enzyme induction in the liver is most likely a completely inappropriate endpoint of toxicity for inhalation exposure.

The importance of airborne PCBs is now understood. Very little is known, however, about the toxicity of these airborne PCBs and the consequences of exposure by inhalation compared to ingestion. Airborne PCBs are lower chlorinated. Our daily exposure to these airborne PCBs may be low under most circumstances, but children playing near Superfund sites in hot summer days, workers moving dried dredging material, or families living unknowingly in buildings with high indoor PCB concentrations, may be exposed to significant amounts of airborne PCBs for extended times.

Cancer in humans (limited, Supplement 7, 1987) The previous IARC review included three occupational epidemiology cohort studies. All three of those studies have been updated (one was expanded) and an additional nine cohort studies have been published. There have also been twelve case-control studies published since the previous IARC review. Many of these used industrial hygienist reviews of participant occupational histories to estimate relative PCB exposure. The strongest studies determined PCB levels in plasma or serum from case and control specimens stored before the cases were diagnosed. No relation was observed between Janus Serum Bank specimen PCB levels and breast cancer diagnoses from the Norwegian Cancer Registry (Ward et al., 2000).

Engel and colleagues found a statistically significant trend of higher odds ratios (OR) for nonHodgkin lymphoma with higher total stored serum PCBs, in three different populations (Engel et al., 2007; Rothman et al., 1997). A significant inverse relation between total PCBs and risk of testicular germ cell tumors was seen in a study linking PCB levels of stored U.S.

Department of Defense (DOD) blood specimens with DOD Medical Surveillance (McGlynn et al., 2009).

Recent papers have reviewed much of the epidemiologic literature (Faroon et al., 2003;

Faroon et al., 2001; Golden and Kimbrough 2009). The lack of congruity in the cohort results may be due to all occupational PCB exposure having been to mixtures of congeners, with the proportion of each congener varying from batch to batch (Hopf et al., 2009). Because there are so many PCB congeners, some co-planar and some not, some estrogenic and some not (Fiedler 1998), it seems plausible that a variety of tumor types could arise from exposure to various congeners, or their metabolites.

Cancer in experimental animals (sufficient, Supplement 7, 1987) Studies indicate that PCB mixtures with a higher chlorine content are more potent in inducing nodular hyperplasia and hepatocarcinomas than mixtures with lower chlorination (Silberhorn et al., 1990), especially in male rodents. In a comprehensive chronic toxicity and carcinogenicity study, the effects of four Aroclor products (1016, 1242, 1254, and 1260) were investigated at multiple dietary concentrations, ranging from 25 to 200 ppm, for 24 months in male and female Sprague–Dawley rats. Statistically significant increases in hepatocellular carcinomas were noted in male rats only for the higher-chlorinated mixture Aroclor 1260, while all four commercial products produced an elevated incidence of hepatocellular carcinomas in female rats. It should be noted that Aroclor 1016 averages only three chlorines per biphenyl. These data indicate that commercial mixtures of chlorinated biphenyls are complete carcinogens, especially in the female rat (Mayes et al., 1998).

Mechanisms of carcinogenicity:

Metabolic Activation of Lower Chlorinated PCBs to Reactive Intermediates It has long been recognized that biphenyl and halogenated biphenyls, particularly the lower halogenated congeners, are hydroxylated in vivo and in vitro (see review by (Safe 1989)).

These hydroxylation reactions are primarily catalyzed by isoforms of cytochrome P-450.

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