REVIEW Chinese Science Bulletin 2004 Vol. 49 No. 3 215ü219
Research progress of the endocrine disrupting activities of polychlorinated biphenyls ZHOU Jingming, QIN Zhanfen, CONG Lin & XU Xiaobai Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences, Beijing 100085, China Correspondence should be addressed to Xu Xiaobai (e-mail: xuxb@ public.bta.net.cn)
Abstract Polychlorinated biphenyls (PCBs) are global persistent organic pollutants. Almost all commercial PCBs mixtures, single PCB congener, and their metabolites possess endocrine disrupting activities. They can disrupt the estrogen/androgen system, thyroid hormone system and other endocrine systems by interfering with the synthesis, transport, storage, metabolism, and feedback regulation of hormones. The newest data related to the endocrine disrupting activities of PCBs and their mechanisms are reviewed and the research perspectives are also discussed. Keywords: polychlorinated biphenyls, endocrine disruptor, thyroid hormone disruptor, environmental estrogen.
chlorinated biphenyls containing 54% of chlorine by weight. They have been used in numerous industrial products such as diluents, dielectric fluids for transformers and capacitors, carbonless copy paper[1]. These compounds are highly lipophilic and persistent and, therefore, are bioconcentrated and biomagnified along the food chain. They have been characterized as global environmental pollutants. PCBs are identified as one of the 12 persistent organic pollutants in the final act of the conference of plenipotentiaries on the Stockholm Convention on Persistent Organic Pollutants in 2001. PCBs are not acute toxic, but subacute and chronic intoxication can lead to skin disorders, behavioral changes, reproductive/developmental abnormalities, and potential carcinogenesis[2]. In recent years, learning deficits and endocrine disrupting actions have been observed in primates exposed perinatally to PCBs. Non-coplanar congener exhibits neurotoxicities via the different mechanisms with coplanar PCB. Clearly, the use of toxic equivalency factors (TEFs) based only on dioxin-like activity is inadequate for estimating the total risk from exposure to PCBs[3]. The following presentation describes experimental results representing the newest information regarding the mechanisms involved in the endocrine disrupting activities of PCBs.
DOI: 10.1360/03wb0147
1 Endocrine disrupting activities of PCBs on the estrogen/androgen system and their mechanisms
Endocrine disruptors are ubiquitous environmental contaminants, which encompass a wide range of substances including natural products, environmental pollutants, pharmaceuticals, and industrial chemicals. It is hypothesized that they may induce a broad spectrum of biochemical and toxic responses at low, environmentally relevant doses. There has been considerable concern regarding the potential adverse effects of endocrine disruptors on the function of development, reproduction, and immune systems of animals and human being. Unfortunately, there are no desirable methods to screen and test endocrine disruptors at present. Polychlorinated biphenyls (PCBs) have been extensively studied and are pervasively recognized endocrine disruptors. Further studying the endocrine disrupting activities of PCBs will contribute to the perfection of screening and testing methods, and the understanding of mechanisms of endocrine disruptions. PCBs are a group of synthetic aromatic compounds which contain a varying number of chlorine atoms substituted on a biphenyl molecule. In theory, there are 209 possible PCB isomers and congeners. Commercial PCBs (e.g. Aroclor, Clophen, Phenoclor, PCB3 and PCB5) are a complex mixture of chlorobiphenyls. All Aroclor formulations are characterized by a four digit number. The first two digits indicate the type of molecule, whereas the last two digits give the percent of chlorine by weight substituted on the molecule. For example, Aroclor 1254 is a mixture of
Bitman and Cecil[4] first reported that ortho-substituted PCB congeners and lower chlorinated Aroclor mixtures such as Aroclor 1221, 1232, 1242 have estrogenic activities. Subsequently, several studies have been reported that hydroxylated metabolites of PCBs competitively bind estrogen receptor (ER), exhibiting agonistic activities[5ü7]. More persistent higher chlorinated hydroxyl-PCBs have been reported to be antiestrogenic[6,8]. PCBs are complex mixtures with a board range of toxic effects. Their endocrine-associated effects mainly associate with their metabolites. The estrogenic activities of PCBs have shown great differences in different test/screen systems. However, this field has made great progress. (ν) Endocrine disrupting activities of PCBs on the estrogen/androgen system. Estrogen is a crucial factor for the development of reproductive system and the sustainment of the second sex characteristics in animals and human being. Moreover, it affects the functions of skeleton and circulating systems. Chemicals that alter synthesis, storage, transport, and metabolism of estrogen may cause estrogen system turbulence. In vitro tests used to screen the estrogenic/androgenic activities of PCBs mainly include estrogen-dependent cell proliferation assay, receptor binding assay, and receptor-mediated reporter gene assay. Estrogen-dependent cell proliferation assay is also called E-screen. This bioassay screens environmental estrogens by comparing the proliferation of estrogen-responsive MCF-7 cells or
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REVIEW T47-D cell between cultures treated with 17E-estradiol and cultures treated with different concentrations of xenobiotics suspected of being estrogenic. The rate of proliferation of MCF-7 cell is elevated significantly by Chinese domestic PCBs (PCB3 and PCB5) at 1011g/mL[9]. Many PCB congeners such as PCB54, 194, 188 and their hydroxylated metabolites all stimulate MCF-7 cell proliferation significantly[10,11]. Primary cultures of rainbow trout hepatocytes retain most of their enzyme systems and synthesize vitellogenin when exposed to 17E-eatradiol or xenoestrogens. This provides a bioassay intermittent between in vitro and in vivo models. Some hydroxylated PCBs such as 4-OH-PCB-50, 4-OH-PCB-30, 4-OH-PCB72, 4-OH-PCB-112 induce vitellogenin synthesis in primary cultured rainbow trout hepatocytes, which are similar to the results obtained from the MCF-7 cell proliferation assay[10]. PCBs have very variable formulations. Metabolism of all the PCB congeners can theoretically yield hundreds of hydroxylated or sulfated metabolites. Determination of the estrogenic activities of each of them would be a huge task, especially since not all the metabolites of each PCB congener have been identified, and the standards are not readily available for all the metabolites[12]. Vakharia et al.[12] developed a system to study in vitro PCB metabolism, in which human liver microsomes produce PCB metabolites in an NADPH-dependent microsomal reaction mixture, and then the metabolites were tested for ER-binding affinity in the human recombinant estrogen receptor-D competitive binding assay. Subsequently, they evaluated the estrogenic activities of metabolites of seven PCBs with different degrees and positions of chlorination using this method. Among the tested PCBs, PCB9, 12, 54, or 156 generated metabolites competed for ER. Moreover, HPLC generated fractions also exhibited ER-binding[13]. Uterotropic activity in immature female or ovariectomized adult female rodents is frequently used as a functional assay for putative estrogenic compounds. Aroclor mixtures, particularly the lower chlorinated ones such as Aroclor 1221, 1232, 1242 are weakly estrogenic in the uterotropic assay, whose potency is 4ü5 orders of magnitude lower than that of 17E-estradiol[4]. Chinese domestic PCB3 and PCB5 have the similar estrogenic potency to that of lower chlorinated Aroclor. Single PCB congener such as PCB54 and 2Ą ,4Ą ,6Ą -tetrachlorinated biphenyl also increase uterine wet weight in rat[11]. Sex determination and sex differentiation of vertebrate have two important models, that is, genetic sex determination (GSD) and environmental sex determination (ESD). The red-eared slide turtle is a species that submits to temperature-dependent sex determination. Exposure to 17E-estradiol may cause sex reversion during the period of sexual differentiation. Bergeron et al.[14] reported that B
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among the 11 tested PCB congeners and their metabolites, 2Ą ,4Ą ,6Ą -trichloro-4-biphenylol and 2Ą ,3Ą ,4Ą ,5Ą -tetrachloro-4-biphenylol significantly reversed sex of turtle at a male-producing temperature. Moreover, when combined, the two compounds synergistically induced sex reversal. Obviously abnormal testes including ovotestes are found in Xenopus laevis after they are exposed to two Chinese domestic PCBs (PCB3 and PCB5) from 5d postfertilization to complete metamorphosis[15]. This result further conforms the estrogenicity of PCBs. Aroclor 1254 interfere with gonadal development with a sex-specific pattern, retarding differentiation of spermatogonia into spermatocytes but accelerating differentiation of oogonia into oocytes in the embryonic chickens[16]. However, Reeder et al.[17] found that at the sites contaminated by PCBs and PCDFs in Illinois, there was a striking sex-ratio reversal in juvenile cricker frogs resulting in a high number of males. Aroclor 1260 also cause abnormal gonads in female rainbow trout, characterized by incomplete or inconsistent development of oocytes at lower dose, and alter the sex ratios at higher dose[18]. There are very different results about the effects of PCBs on gonadal development due to the different experiment designs, animals and PCBs. In general, it can be concluded that PCBs interfere with sex determination and sex differentiation, induce ovotestes and sex reversal in animals. Some chemicals may exhibit antiandrogenic effects by interfering with the binding of androgen to androgen receptors. Aroclor 1016, 1221, 1232, 1242, 1248, 1254, and 1260 act antagonistically at concentrations that do not affect cell viability in human androgen receptor reporter gene assay[19]. Some individual PCB congeners such as PCB49, 66, 74, 105, and 118 completely antagonize the stimulation by 5D-dihydrotestosterone (DHT), and PCB138, 153, 156 are less effective antagonists[19,20]. A proportionally representative mixture of PCBs that detected in human milk also cause the DHT-mediated activation of luciferase activity to be reduced by more than 50%[19]. 2,3,7,8-tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD) induce cytochrome P450 monooxygenase 1A1, 1A2 via aryl hydrocarbon receptor (AhR), which involved in the metabolism of estrogens. Coplanar PCBs bind to the AhR and exhibit activities similar to that of TCDD. Monoortho-substituted analogs of the coplanar PCBs also exhibit AhR agonist activity[1]. It is hypothesized that coplanar and monoortho-substituted PCBs may exhibit antiestrogenic activity through the same mechanisms as TCDD. It has been proven that PCB77, 126, 39 have antiestrogenic activities in MCF-7 cell proliferation assay and/or uterotropic assay. Several hydroxylated PCBs identified in human serum exhibit minimal binding to the rat uterine and do not induce proliferation of MCF-7 cell at concentrations ranging from 105 to 108 mol/L. Moreover, when cotreated with 17E-estradiol, hydroxylated PCBs inhibited B
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REVIEW 17E-estradiol-induced cell proliferation[8]. However, Kramer et al.[21] reproted that among the 13 OH-PCBs tested in an in vitro bioassay, 11 OH-PCBs are antiestrogenic at three different concentrations of 17E-estradiol. But this “antiestrogenicity” is related to their effects on cell viability and, therefore, it cannot be described as exhibiting “hormone disruption” solely by an estrogen receptor mediated mechanism. Based on this result, it may be concluded that OH-PCBs would not exhibit detectible antiestrogenic activity at physiologically relevant 17Eestradiol concentrations. PCBs have long been known to be estrogenic, but in vitro estrogen receptor binding is not strong except for some lower chlorinated hydroxyl metabolites. More persistent higher chlorinated OH-PCBs have been reported to be antiestrogenic. Coplanar AhR agonists are also generally considered to be antiestrogenic[22]. Korach and coworkers[5] first reported that hydroxyl-PCBs competitively bound mouse estrogen receptor and two of the most active ER agonists in their study were 2Ą ,4Ą ,6Ą trichloro-4-biphenylol and 2Ą ,3Ą ,4Ą ,5Ą -tetrachloro-4-biphenylol. Several studies suggest that hydroxylation is important for PCBs interaction with ER, and congeners containing a single para-hydroxy group on one of the two biphenyl rings have the most estrogenic activities. Conner et al.[7] performed a systematic study on the structureestrogenicity relationships for hydroxyl-PCBs, and then concluded the structure-estrogenicity relationships for hydroxyl-PCBs were complex and response-specific. (ξ) Mechanisms of estrogenic/androgenic activities of PCBs. Estrogens are eliminated from body by metabolic conversion to hormonally inactive (or less active) water-soluble metabolites. The metabolic disposition of estrogens includes oxidative metabolism, and conjugative metabolism by glucuronidation, sulfonation and/or O-methylation[23]. There is only limited information available related to the effects of PCBs on estrogen metabolism. The main enzyme system involved in estrogen metabolism is the cytochrome P450 system. 2,3,7,8-TCDD and related compound affect aromatase, thus interfering with the conversion of testosterone to estradiol[24]. Coplanar PCB126 increase the 7D-hydroxylation but suppress the 2D-, 6E-, and 16D-hydroxylation and 5D-reduction of progesterone and testosterone in liver microsomes[25]. Estrogen sulfotransferase (EST) is a cytosolic enzyme that catalyzes the sulfoconjugation and inactivation of estrogens. Sulfation by estrogen sulfotransferase is an important pathway for 17E-estradiol inactivation. PCB77 do not affect EST activity even at the highest concentration (1000 nM). However, hydroxylated PCBs induce strong inhibitory activity that is dependent on the positions of the substituents in the phenyl ring[26]. By inhibiting the formation of inactive 17E-estradiol sulfate, OH-PCB can increase 17E-estradiol bioavailability in target tissue, thereby exChinese Science Bulletin Vol. 49 No. 3
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erting an indirect estrogenic effect. This result provides a new and receptor-independent paradigm to explain the estrogenic activity of PCBs. ERs are ligand-activated transcription factors and members of the nuclear receptor superfamily. ERs bind estrogenic steroids and structurally diverse compounds with different affinities. The parent PCBs themselves are not effective to bind the estrogen receptor. Some hydroxyl-PCBs competitively bind to the estrogen receptors[5,7,12,13]. PCB phenol-ER complexes translocate into nucleus, bind to ER response elements, and elicit the estrogenic activities. PCBs have been also shown to be antiandrogenic by interfering with the binding of androgen to the androgen receptor[19,20]. Aroclor 1254 may impair GnRH synthesis in the preoptic anterior hypothalamic area (POAH) when Atlantic croaker exposed to Aroclor 1254 for 30 d during the early-recrudescence phase of the gonadal cycle. The GnRH content in the POAH, number of pituitary GnRH receptor, and LH secretion in the PCB-exposed group are comparable to those in the early-recrudescence phase. This impairment may be due to a direct action of PCB on GnRH neurons and/or indirectly via interference with other neurotransmitter pathways that modulate GnRH function[27]. The hypothalamic GT1-7 cells can synthesize and secrete GnRH, and have many similarities to GnRH neurons in vivo. Aroclor 1221 are stimulatory to GnRH gene expression, however, Aroclor 1254 inhibit GnRH nuclear mRNA levels at high doses, and stimulate GnRH mRNA at low doses. These results demonstrate a novel mechanism for effects of the two PCBs directly on GnRH gene expression, and indicate a hypothalamic level for endocrine disruption of PCBs. 2 Endocrine disrupting activities of PCBs on the thyroid hormone system and their mechanisms Thyroid hormones (THs) are essential for normal body metabolism, growth, and development including reproduction, maturation, and aging. There are many examples of pharmaceutical, environmental, and naturally occurring chemicals that alter one or more of the processes involved in the synthesis, storage, release, transport, and metabolism of THs. In addition to environmental estrogen, thyroid hormone disruptors have become the most important endocrine disruptors. (ν) Endocrine disrupting activities of PCBs on the thyroid hormone system. The effects of PCBs on the thyroid hormone system are more potent than that on the estrogen system. Almost all the commercial PCBs mixtures and PCB congeners can interfere with the thyroid hormone homeostasis. Most information on thyroid hormone-related effects of PCBs stem from relatively short-term single-dose exposure studies of experimental animals exposed to either mixtures or PCB congeners. Exposure to PCBs mixtures (Aroclor 1254, 1260, Clophen 217
REVIEW A50) or to single PCB congeners (PCB77, 105, 118, 126, 169, and 153) result in marked reductions in plasma levels of T4, whereas T3 levels are only marginally affected. Reductions in plasma thyroid hormone levels are observed 24ü48 h after a single intraperitoneal injection and remain low for prolonged periods of time (several days to months), depending on the half-life of the PCB congener[25]. (ξ) Mechanisms of thyroid hormone disrupting activities of PCBs. Exposure to PCBs may result in thyroid gland histological and ultrastructural changes. In rats administered PCBs there is an enlargement of the thyroid gland and the thyroid follicular cells accumulate large numbers of colloid droplets and large abnormally shaped lysosomes[28,29]. Ultrastructurally, PCB-induced changes in thyroid gland include an increased development of endoplasmic reticulum, vacuolization of mitochondria and a decrease in the colloid droplet-lysosome interaction necessary for the secretion of thyroid hormones[28]. The ultrastructural lesions produced by PCBs are distinct from stimulation by exogenous thyrotropin, pituitary suppression by exogenous thyroxine, and stimulation of follicular cells by feeding a low iodine diet. These studies show evidences of possible direct effects of PCB on the thyroid gland[30]. Coplanar PCBs and TCDD cause reductions in thyroid hormones mainly by binding to AhR and induce hepatic UDP-glucuronosyl transferases (UGTs) that result in increased T4 biliary excretion[31]. There are at least two or possibly three UGT isozymes involved in thyroid hormone glucuronidation. T4 glucuronidation is markedly induced in the livers of rats exposed to PCB77, 169, 126, and 156. The nonplanar di-ortho congener PCB153 and PCBs mixtures are also capable of inducing T4 glucuronidation[25]. PCBs also affect thyroid hormone deiodinases (IDs) as well as sulfotransferases (SULT). Exposure of rats to PCB77 result in an inhibition of hepatic ID-1 that converts T4 into T3 as well as reverse T3 activity[32]. OH-PCBs are also potent inhibitors of thyroid hormone sulfation, competitively inhibiting the SULT enzyme[33]. Inhibition of ID activity or SULT activities may compromise the availability of intracellular T3. Disruption of TH transport is one of the key mechanisms by which PCBs alter TH homeostasis. PCB metabolites with hydroxyl groups on meta or para positions have structural resemblance to T4 and affinity for transthyretin (TTR), a thyroid hormone plasma transport protein. Thus, T4 is competitively displaced by PCB metabolites, rendering it vulnerable to metabolism and elimination[34]. The binding of OH-PCBs to TTR may be involved in the rapid and facilitated transfer of OH-PCBs instead of the natural hormone T4 across the placental and blood-brain barriers, leading to relatively high levels of OH-PCBs accumulated in the fetus, particularly in the fetal brain[35,36]. It should be noted that some mammals including humans have a sec218
ond binding protein, thyroxine-binding protein, and no competitive interaction on thyroxine-binding globulin takes place with OH-PCBs, which may indicate a lower impact of OH-PCBs on plasma T4 levels in those species possessing thyroxine-binding globulin. Serum thyroid hormones are regulated by the hypothalamo-pituitary-thyroid (HPT) axis. Effects of PCB exposure on thyroid hormone feedback regulation are equivocal. Some reports indicate an increased TSH concomitant with a low T4 level after exposure to PCBs. However, other reports have not observed any effect on TSH despite severely depressed T4 levels. Khan et al.[37,38] reported that ortho-PCB congeners PCB95 and PCB 101 may interfere with the HPT-axis by causing a subnormal response of the pituitary and thyroid to TRH stimulation. Biologically active TH mediate its signal by binding to thyroid hormone receptors (TRs). TRs are ligandinduced transcription factors that regulate target genes. Due to the similarity of the structures of TH and PCBs, the competitive binding of PCBs to TR has been proposed[39]. However, the binding affinities of some hydroxylated PCBs are less than 1/10000 of T4. Low dose of PCBs can potentially interfere with TR-mediated transactivation by influencing on TR/coactivator complex[40]. In addition to estrogen/androgen and thyroid hormone systems, PCBs also disrupt the retinoid and adrenal system. The effects of PCBs involved in the whole neuro-immunoendocrine system. Further studies should be conducted for comprehensively evaluating the endocrine disrupting activities of PCBs. 3
Prospects
Although their manufacture has been banned for about 30 years, PCBs are routinely detectable pollutants in air, water, sediments, fish, wildlife, and human tissues[1]. There is no desirable method to dispose PCBs except for high temperature incineration. It is possible that the sealed PCBs may enter environment in the future. The risk that PCBs imposed to animals, human, even the whole ecosystems still exists. It is necessary to further investigate the PCBs levels in all environmental mediums and to reevaluate the environmental risk of PCBs. Much work should be focused on developing the analysis methods of PCBs, carrying out epidemic investigations in polluted areas, exploring the molecular mechanisms of the toxicity of PCBs, conducting quantitative structure-activity relationships (QSAR) studies, and developing animal models to study the dose-effects relations at low, environmental relevant doses. Acknowledgements This work was supported by the Chinese Academy of Sciences (Grant No. KZCX-414) and the “863” Program (Grant No.2001AA640601).
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