Development of TEFs for PCB congeners by using an alternative biomarker — Thyroid hormone levels

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Abstract

Polychlorinated biphenyls (PCBs) are ubiquitous toxic contaminants. Health risk assessment for this class of chemicals is complex: the current toxic equivalency factor (TEF) method covers dioxin-like (DL-) PCBs, dibenzofurans, and dioxins, but excludes non-DL-PCBs. To address this deficiency, we evaluated published data for several PCB congeners to determine common biomarkers of effect. We found that the most sensitive biomarkers for DL-non-ortho-PCB 77 and PCB 126 are liver enzyme (e.g., ethoxyresorufin-O-deethylase, EROD) induction, circulating thyroxine (T4) decrease, and brain dopamine (DA) elevation. For DL-ortho-PCB 118 and non-DL-ortho-PCB 28 and PCB 153, the most sensitive biomarkers are brain DA decrease and circulating T4 decrease. The only consistent biomarker for both DL- and non-DL-PCBs is circulating T4 decrease. The calculated TEF-TH, based on the effective dose to decrease T4 by 30% (ED30) with reference to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), is identical to both TEF-WHO98 and TEF-WHO05 for TCDD and DL-PCBs (correlation coefficients are r = 1.00, P < 0.001; and r = 0.99, P < 0.001, respectively). We conclude that T4 decrease is a prospective biomarker for generating a new TEF scheme which includes some non-DL-congeners. The new TEF-TH parallels the TEF-WHO for DL-PCBs and, most importantly, is useful for non-DL-PCBs in risk assessment to address thyroid endocrine disruption and potentially the neurotoxic effects of PCBs.

Introduction

Polychlorinated biphenyls (PCBs) are a family of theoretically 209 congeners, consisting of two benzene rings and 1–10 chlorine atoms (Kafafi et al., 1993). PCBs range from light oily fluids to heavy greasy substances and are colorless to light yellow in color with no known smell or taste. Because of PCBs’ remarkable insulating capacity and their flame retardant nature, PCBs gained widespread use in transformers and other electrical equipment in the U.S. from 1929 to 1977. However, PCB technical mixtures are composed of a smaller suite of congeners, and only about 80–100 PCB congeners are of actual environmental relevance (Maervoet et al., 2004). There are different classifications of PCBs. The United States Environmental Protection Agency (USEPA, 2003) defined all 20 non-ortho-PCBs and 48 mono-ortho-PCBs, based on their planar structural configuration (which is related to their dioxin-like (DL-) toxicity), as coplanar PCBs. The World Health Organization (WHO, 2003) identified 3 non-ortho-PCBs 77, 126, and 169, 2 mono-ortho-PCB 105 and PCB 118, and 6 additional PCBs 28, 52, 101, 138, 153, and 180 as particularly important PCB congeners, based on their presence in human breast milk (an indicator of human exposures).

PCBs are routinely found in samples of human and animal tissues. Health risk assessment for this class of chemicals is complicated by the fact that humans are exposed to mixtures of many PCB congeners in differing proportions. Individual PCB congeners exhibit different physicochemical properties and biological activities that result in different environmental distributions. In addition, PCBs’ different toxic mechanisms complicate risk assessment. PCB congeners’ potencies for several distinct toxic effects vary according to structures and chlorine-substitution on the biphenyl rings. In general, most coplanar congeners cause a suite of DL-responses, including carcinogenicity, associated with activation of the aryl hydrocarbon (Ah-) receptor. In contrast, nonplanar PCBs elicit a diverse spectrum of non-Ah-receptor-mediated toxic responses in humans and animals, including neurotoxic (Kenet et al., 2007, Seegal, 1996), carcinogenic (Carpenter, 2006, Knerr and Schrenk, 2006), and endocrine effects (Crofton et al., 2005, Koopman-Esseboom et al., 1994, Ness et al., 1993).

Risk assessment for DL-PCBs and other DL-compounds uses a toxic equivalency factor (TEF) approach, relating the potency of individual congeners to that of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) (Van den Berg et al., 2006). The TEF method is based on congener binding and activation of Ah-receptor-mediated enzyme activities. TEFs are used to calculate toxic equivalent (TEQ) that is the sum of all individual congener’s TEF multiplied by each congener’s concentration in the mixture. Unfortunately, no such unifying methodology currently exists for the non-DL-PCBs and few pure congeners have been extensively studied. To include non-DL-PCBs in the risk assessment and, at the same time, to supplement the existing WHO TEF scheme for the DL-effects, we believe that an additional TEF method to cover other effects of both DL- and non-DL-PCBs is warranted.

The majority of adult human exposure to environmental PCBs is via food products, although other routes (i.e., air, water, and soil contact) may also contribute. Adults exposed to higher concentrations of PCBs and thermal breakdown products from contaminated rice oil in the Yusho and Yu-Cheng incidents developed chloracne, fatigue, nausea, and liver disorders, and some newborns of exposed mothers exhibited developmental abnormalities (Schantz, 1996a). Occupational and environmental exposures to PCBs have drawn attention primarily because of the carcinogenicity of PCBs (Ahrens et al., 2007, Brown and Jones, 1981, Priha et al., 2005, Prince et al., 2006, Ruder et al., 2006). A Greenlandic Inuit study found that blood plasma PCBs are inversely associated with global DNA methylation — an epigenetic mechanism that has been associated with cancer initiation and progression (Rusiecki et al., 2008). Although there is evidence for increased cancer risk/mortality from both occupational and environmental PCB exposures (De Roos et al., 2005, Demers et al., 2002, Nelson, 2005, Salehi et al., 2008), PCBs are classified as “probable human carcinogens” by the WHO and “class 2B” by the International Agency for Research on Cancer (IARC), based on insufficient human evidence (Carpenter, 2006), but sufficient evidence of carcinogenicity in animals. Both DL- and non-DL-PCB congeners can promote cancers (Knerr and Schrenk, 2006).

The developing fetus and infants are especially vulnerable to environmental pollutants (Landrigan et al., 2004, Miller et al., 2002). Infants are at risk for neurotoxicity from PCB exposures in utero and during breast feeding (Carpenter, 2006). Even low levels of PCBs transferred to the fetus across the placenta and to neonates via breast milk may induce long-lasting neurological damage (Jacobson and Jacobson, 1996). The potential neurotoxicity of PCBs was first recognized in 1968 following the Yusho incident (Schantz, 1996a). Subsequently, several epidemiological studies in different parts of the world reported neurotoxic effects of PCB exposure. The mothers in these studies generally exhibited no signs of PCB toxicity. Lower growth rate in children was associated with prenatal PCB exposure in a Dutch cohort (Patandin et al., 1998) and was reported in some of the infants born to mothers who consumed contaminated Great Lakes fish in the U.S. (Darvill et al., 2000). Some consistent observations at birth have been motor immaturity and hyporeflexia and lower psychomotor scores between 6 months and 2 years old (Faroon et al., 2000). Poorer cognitive functioning was noted in preschool children (Stewart et al., 2003). Most of the available epidemiological data to date have reported associations between prenatal PCB exposure and decreases in measures of cognitive functioning in infants and children (Darvill et al., 2000, Jacobson and Jacobson, 1996). Even in the Oswego cohort, where the levels of exposure are significantly lower than that in the earlier studies, negative impacts on cognitive functioning are seen (Darvill et al., 2000).

PCB-induced neurotoxicity may relate to its ability to decrease circulating thyroid hormone (TH). Sufficient TH in early pregnancy is critical for normal brain development, and immature animals and children are differentially sensitive to TH disruption during critical windows of development (Hansen, 1998). In the first 12 weeks of gestation, human fetuses depend entirely on TH of maternal origin as their thyroid is not functional (Calvo et al., 2002). The fetal thyroid assumes an increasing role in producing TH as gestation progresses (Zoeller and Rovet, 2004). Human TH insufficiency during fetal development mainly affects cortical development, while postnatal hypothyroidism exerts effects on cerebellar development (Zoeller and Rovet, 2004). Studies found that levels of free thyroxine (FT4) and the presence of circulating antibodies for thyroid peroxidase were strong predictors of infant mental development and children’s intelligence quotient (IQ) (Pop et al., 1995). Moreover, PCBs have been shown to have neurotoxic effects and to alter thyroid function during critical periods of TH-dependent brain development (Porterfield, 2000). A decreased fetal or maternal TH supply in pregnancy is associated with infants’ poorer attention and altered rates of information processing (Mirabella et al., 2000). TH deficiency early in pregnancy causes problems in visual attention, visual processing (i.e., acuity and strabismus) and gross motor skills in offspring. TH deficiency later in pregnancy increases the risk of subnormal visual (i.e., contrast sensitivity) and visuospatial skills in the child, as well as slower response speeds and fine motor deficits, while postnatal TH deficiency predominantly affects language and memory skills (Zoeller and Rovet, 2004). Newborn infants with severe hypothyroidism had an IQ reduction of about 4 points if untreated during the first 6–12 months (Burrow et al., 1994, Gyamfi et al., 2009). Treatment with TH soon after birth can recover some function, although these children are still at risk for mild learning disabilities (Porterfield, 2000). One recent report revealed that prenatal PCB exposures were associated with reduced total thyroxine (T4) and FT4 levels of infants in the U.S. (Herbstman et al., 2008). Thus, there is increasing concern for prenatal and postnatal PCB exposures in regard to both thyrotoxic and neurotoxic effects (Arisawa et al., 2005, Crofton et al., 2000, Longnecker et al., 2000). Therefore, as a step toward establishing a new TEF method that will cover both DL- and non-DL-PCB congeners, we asked whether thyrotoxic or neurotoxic biomarkers can be used in PCB congener-specific risk assessment.

Section snippets

Selection of PCB congeners for TEF-TH

In order to select the most appropriate PCB congeners for which to develop TEFs based on TH effects (TEF-TH), we evaluated the literature on exposure as well as toxicity. We examined literature on the presence of PCBs in human milk and dairy products, as these are important sources of exposure for infants. Studies of the concentration of different PCBs in pooled human milk samples from seven different countries found that non-ortho-PCBs 77, 126, 169, mono-ortho-PCBs 28, 105, 118, di-ortho-PCBs

Comparison of liver, thyroid, and brain toxicity in rats

We wanted to determine whether a decrease in circulating T4 is a good biomarker to correlate with toxic responses for both DL- and non-DL-PCB congeners. Studies in rats have demonstrated that most PCB congeners, such as non-ortho-PCBs 77, 126, and 169 (Crofton et al., 2005, Morse et al., 1993, Seo et al., 1995, Van Birgelen et al., 1995), mono-ortho-PCB 28 and PCB 118 (Chu et al., 1995, Ness et al., 1993), di-ortho-PCB 101 and PCB 153 (Khan et al., 2002, Ness et al., 1993), and tri-ortho-PCB 95

Discussion

In this study, we found that T4 decrease is an important alternative biomarker to develop a TEF method that will cover both DL- and non-DL-PCB congeners and more specifically assess thyroid endocrine disruption and neurotoxicity. The newly developed TEF-TH correlates to the WHO TEFs very well for DL-PCBs (Fig. 6) and, most importantly, it also covers non-DL-PCBs (Table 2). This new TEF-TH method takes a step forward to address the deficiency in the current methods for assessing risk of PCBs

Acknowledgments

We thank Drs. Kevin M. Crofton (U.S. EPA), Ling-Hong Li, Amy Arcus-Arth, and James Collins (Cal/EPA) for their critical reading of the document and helpful comments.

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    The views expressed in this article are those of the authors and do not necessarily reflect those of the Office of Environmental Health Hazard Assessment, the California Environmental Protection Agency, or the State of California. The authors declare they have no competing financial interests.

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