Estrogenic Endocrine Disruptors: Molecular Characteristics and Human Impacts

2.31 Estrogenic Endocrine Disruptors: Molecular Characteristics and Human Impacts S Safe, I Jutooru, and G Chadalapaka, Texas A&M University, TX, USA ª 2010 Elsevier Ltd. All rights reserved.

2.31.1 2.31.2 2.31.3 2.31.3.1 2.31.3.2 2.31.3.2.1 2.31.3.2.2 2.31.3.3 2.31.4 2.31.4.1 2.31.4.2 2.31.4.3 2.31.4.4 2.31.5 2.31.6 2.31.7 References

Introduction Endocrine Disruptors and Hormone-Dependent Diseases Endocrine Disruptors: Chemicals of Concern Introduction Conflicting Results of Laboratory Animal Studies DES studies BPA studies BPA: Is It a Prototypical Endocrine Disruptor? Xenoestrogens and Phytoestrogens as SERMs: Implications for Risk Assessment Multiple Pathways for ER Activation Factors that Influence ER Activation SERMs Differentially Activate Wild-type and Variant ER SERMs Differentially Activate Wild-type and Variant ER/Sp EDs and their Effects on Male Reproductive Tract Problems Organochlorine EDs and Breast Cancer Conclusions

Abbreviations ADHD AP BPA DES DHT ED EE EPA ERE NIEHS

attention deficit hyperactivity disorder alkylphenol bisphenol A diethylstilbestrol dihydrotestosterone endocrine disruptor ethinylestradiol Environmental Protection Agency estrogen responsive element National Institute for Environmental Health Sciences

2.31.1 Introduction Endocrine systems are critical determinants for development and growth of multiple tissues and organs and for maintaining homeostasis. Endocrine pathways are intimately associated with intercellular communication and involve secretion of a molecule from one cell type that influences other cells in the same tissue/organ or at distal sites. The types of secreted molecules are highly variable and can

OC PCB PCDD PCDF PR SERM TDS VEGFR

609 610 611 611 612 612 612 613 613 613 614 615 616 618 619 619 619

organochlorine polychlorinated biphenyl polychlorinated dibenzo-p-dioxin polychlorinated dibenzofuran progesterone receptor selective estrogen receptor modulator testicular dysgenesis syndrome vascular endothelial growth factor receptor

include polypeptides or low molecular weight compounds such as the steroid hormones 17 -estradiol (E2) and dihydrotestosterone (DHT) which are major estrogenic and androgenic hormones, respectively. Recognition of secreted molecules may involve both cell surface or cell membrane receptor or acceptor sites or intracellular receptors that are expressed at various locations within the cell including the nucleus. The neuroendocrine system is an example of a critical endocrine pathway which is 609

610 Alterations in Cell Signaling

initiated by signals originating in the central nervous system to activate the hypothalamus. This results in secretion of a series of hypothalamic releasing factors which in turn act on both the anterior and posterior pituitary glands. The pituitary glands then release polypeptide hormones such as corticotropin, follicle stimulating hormone, somatropins, luteinizing hormone, prolactin, thyrotropin, oxytocin, and vasopressin which in turn activate specific tissues of secondary intracellular targets such as the adrenal and thyroid glands, ovaries, and testes. Subsequent activation of secondary targets results in stimulation of biochemical and physiological responses in multiple downstream tissues. Not surprisingly, defects in the endocrine system can result in multiple diseases including diabetes, Graves disease, and many other metabolic disorders which are associated with mutations, deletions, or overexpression of hormones, hormone receptors, and their associated downstream biochemical partners. One of the paradoxes of endogenous hormones such as E2 is that even in women with apparent normal estrogenic function, there is strong evidence from epidemiology studies that lifetime exposure to E2 and other estrogenic hormones is a known risk factor for breast cancer (Lambe et al. 1996; Meeske et al. 2004; Russo et al. 1992; Talamini et al. 1996). Thus, early menarche, late menopause, age at first birth, and nulliparity, which enhance lifetime exposure to unopposed estrogens, increase the risk for breast cancer in women.

2.31.2 Endocrine Disruptors and Hormone-Dependent Diseases The role of environmental chemicals or other xenobiotics including drugs as causal or contributing agents in endocrine-related human diseases was proposed by Theo Coburn and other scientists in the early 1990s (Colborn et al. 1993; Davis and Bradlow 1995; Davis et al. 1993; Giwercman et al. 1993; Sharpe and Skakkebaek 1993). The initial hypotheses suggested that endocrine disruptors (EDs) and particularly those ED compounds that exhibit estrogenic activity (xenoestrogens) were responsible for the widespread decrease in male reproductive capacity and the increased incidence of breast cancer in women. It was further emphasized that the timing of exposure to these compounds was also a critical factor, and in utero or early postnatal exposures were the most important exposure windows that result in subsequent endocrine dysfunction. The

initial support for this hypothesis was provided by wildlife, laboratory animal, and human studies. Evaluation of wildlife studies in regions contaminated with organochlorine (OC) chemicals such as polychlorinated biphenyls (PCBs), polychlorinated dibenzo-p-dioxins (PCDDs), polychlorinated dibenzofurans (PCDFs), and pesticides showed that some of their reproductive problems and increased tumors could have been contaminant-related. This is supported, in part, by recoveries of wildlife and fish populations in some areas where contaminant levels have decreased due to a combination of banning or restricting the use of OC chemicals and time-dependent biodegradation or lack of bioavailability (e.g., due to sedimentation) (Tremblay and Gilman 1995). Several estrogenic industrial compounds such as the alkylphenols (APs) have been found in river systems and sewage outflows, and it was proposed that the estrogenic activity of these compounds contributed to the ‘feminization’ of male fish in British rivers and other aquatic environments (Harries et al. 1997; Purdom et al. 1994; Sumpter and Jobling 1995). Subsequent studies by Sumpter and colleagues in the United Kingdom show that the major estrogenic compounds in these rivers were the endogenous estrogenic hormones E2 and estrone and the birth control drug, 17-ethinylestradiol (EE) (Desbrow et al. 1998; Johnson et al. 2006; Routledge et al. 1998). Thus, it appeared that excreted hormones in household sewage and run-off from animal feces/urine were the major estrogenic contaminants in rivers with EE also contributing to the overall activity. Early evidence that endocrine disrupting compounds were linked to a global decrease in male reproductive tract problems and an increased incidence in breast cancer in women was initially dependent on a select number of studies (Carlsen et al. 1992; Falck et al. 1992; Wolff et al. 1993). An important underpinning for the effects of EDCs on males and females was the disastrous experience with the estrogenic drug diethylstilbestrol (DES) (Bibbo et al. 1977; Gill et al. 1979; Giusti et al. 1995; Henderson et al. 1976; Newbold 1995, 2008; Stillman 1982). DES (Figure 1) is a highly potent estrogenic drug given to women during pregnancy until it was discovered that in utero exposure to this compound had profound adverse effects on male and female offspring. DES sons and daughters exhibited a host of reproductive tract problems including deformed genitalia in males and a high incidence of vaginal adenocarcinoma in females. Results obtained for DES, coupled with the observation in a

Estrogenic Endocrine Disruptors: Molecular Characteristics and Human Impacts

611

Figure 1 Structures of 17 -estradiol, DES, tamoxifen, and the organochlorine (OC) compounds PCBs and DDE.

meta-analysis of sperm count studies in males from fertility clinics showing an overall global decrease in sperm counts from 1950 to 1990 (Carlsen et al. 1992), led to the hypothesis that ‘xenoestrogens’ and possibly dietary estrogens contributed to the sperm count decrease. It was also hypothesized that the environmental etiology of these effects was strengthened with the report showing that 1,1-dichloro-2,2-bis(pchlorophenyl)ethylene (DDE), a major environmental contaminant in humans, was also an antiandrogen (Kelce et al. 1995; Sharpe 1995) (Figure 1). This was an important observation since antiandrogens also induce a host of reproductive tract problems in laboratory animal models. Support for the role of xenoestrogens in breast cancer came from small case-control studies in Connecticut and New York showing that PCB and DDE levels were increased in tissue or serum from breast cancer patients compared to controls (Falck et al. 1992; Wolff et al. 1993). All of these studies were published from 1990 to 1995, and not surprisingly, these results and their implications generated public, media, regulatory, and scientific concerns which were addressed, in part, by initiation of human, laboratory animal, and mechanistic studies to test the validity of the ED hypothesis.

2.31.3 Endocrine Disruptors: Chemicals of Concern 2.31.3.1

Introduction

The ED hypothesis resulted in a legislative mandate to the Environmental Protection Agency (EPA) to develop and validate screening and testing procedures for endocrine disrupting chemicals that

modulate the estrogen, androgen, and thyroid hormone receptors. This legislation also resulted in several scientific initiatives by the EPA, the National Institute for Environmental Health Sciences (NIEHS), and other funding agencies to not only identify EDs of concern but also to undertake laboratory animal and mechanistic studies to understand their action. Prior to the mid-1990s, the estrogenic activity of several environmental chemicals had already been identified and these included some PCBs, kepone, animal feed supplements such as zearalenone, o,p-DDT, methoxychlor and its metabolites, phenol red, other OC insecticides, bisphenol A (BPA), and APs (Berthois et al. 1986; Bitman and Cecil 1970; Bitman et al. 1968; Ecobichon and MacKenzie 1974; Hammond et al. 1979; Korach et al. 1988; Krishnan et al. 1993; Robinson et al. 1984; Soto et al. 1991; Tullner 1961; Welch et al. 1969; White et al. 1994). The more recent characterization of BPA and APs as estrogenic compounds was particularly significant since these are high-volume industrial chemicals with potential for both human and environmental exposures. Figure 2 lists some of the EDs of concern with respect to human exposures and these include APs, BPA, organochlorinated insecticides, PCBs, and a variety of compounds in personal care products such as alkylhydroxybenzoates (parabens) and 4-methylbenzylidene camphor (UV filters). The potential adverse human impacts of these compounds are controversial and this is due to several factors. First, human exposure to synthetic or xenoestrogenic compounds must be evaluated in terms of their overall exposure to estrogenic and antiestrogenic compounds in the diet which includes relatively high exposures to estrogenic and

612 Alterations in Cell Signaling

Figure 2 Structures of some endocrine disruptors (EDs) of concern. Octylphenol (OP) is similar to NP but contains a C8H17 side chain.

antiestrogenic flavonoids and other phytochemicals. For example, using MCF-7 breast cancer cells as a model for estrogenic activity, the overall estrogen equivalents from xenoestrogens in the diet is significantly lower than the corresponding activity of a glass of red wine or vegetables that express estrogenic phytochemicals (Gaido et al. 1998). This comparison of estrogenic potency also holds for serum levels of endogenous steroids, estrogenic phytochemicals, and synthetic xenoestrogens (Safe 2000; Safe et al. 1997). There are at least two additional issues which lend to the uncertainty regarding potential adverse human health effects due to exposure to EDs. Several laboratory animal studies report highly conflicting results regarding the effects of EDs such as BPA. There is also evidence that the potential effects of estrogenic EDs are tissue-selective, suggesting that these compounds are selective estrogen receptor modulators (SERMs). Both of these issues and their significance are discussed below (Gray et al. 2004; Safe 2005; Vom Saal et al. 2007).

2.31.3.2 Conflicting Results of Laboratory Animal Studies 2.31.3.2.1

DES studies The ED hypothesis suggests that the potential adverse impacts of xenoestrogens and possible antiandrogens (such as DDE) are due to in utero or early postnatal exposures during critical susceptible periods of male and female reproductive tract development. There are numerous examples where

in utero exposure to compounds such as 2,3,7,8-tetrachlorodibenzo-p-dioxin, other Ah-receptor agonists, and ED compounds, such as DES, imprints various genes and this results in significant effects in the adult animals (Bibbo et al. 1977; Gill et al. 1979; Giusti et al. 1995; Henderson et al. 1976; Mably et al. 1992; Newbold 1995, 2008; Stillman 1982). An extensive study on the effects of in utero exposure of mice and humans to the estrogenic drug DES has revealed that many of the same responses have been observed in both species (Bibbo et al. 1977; Gill et al. 1979; Giusti et al. 1995; Henderson et al. 1976; Newbold 1995, 2008; Stillman 1982). Individuals exposed to DES in utero exhibit a diverse spectrum of reproductive tract problems and there is some evidence that some of these effects may even be observed in their offspring. In addition, there is also evidence that offspring of mice exposed in utero exhibit increased tumors. It is clear that DES demonstrates the potential for inducing adverse effects in human populations; however, it should also be noted that in ER binding and ERdependent transactivation assays DES is a potent estrogen with activity equal to or greater than the endogenous hormone E2 (Kuiper et al. 1998). Moreover, in many cases, the pharmacologic doses of E2 given to pregnant women were exceedingly high. 2.31.3.2.2

BPA studies The major controversy with respect to EDs of concern is associated with BPA. This compound is used as a monomer in the production of resins and polycarbonate plastic products which have wide use in

Estrogenic Endocrine Disruptors: Molecular Characteristics and Human Impacts

consumer and industrial materials. It is estimated that over six billion pounds of BPA is used each year and parts per billion levels of BPA are detected in the urine of individuals living in developed countries (Calafat et al. 2008). The persistence of BPA is somewhat surprising due to the phenol groups which should undergo phase II metabolism to form conjugates that are readily excreted. Nevertheless, BPA is a high-volume chemical which is now routinely detected in humans and there is concern by some groups that BPA may negatively impact various aspects of human health which include ‘‘increases in abnormal penile/urethra development in males, early sexual maturation in females, an increase in neurobehavioral problems such as attention deficit hyperactivity disorder (ADHD) and autism, an increase in childhood and adult obesity and type 2 diabetes, a regional decrease in sperm count, and an increase in hormonally mediated cancers, such as prostate and breast cancers’’ (Vom Saal et al. 2007). In contrast, a study carried out by the Harvard Center for Risk Analysis to evaluate the ‘low dose’ effects data for BPA concluded the following: ‘‘In the case of BPA, the evidence considered by the panel suggests that the weight of the evidence for low-dose effects is very weak. Studies are conflicting, the effects are subtle with questionable functional importance even if real, and there are conflicting data as to the proposed mode of action (i.e., whether BPA acts as an estrogen)’’ (Gray et al. 2004). 2.31.3.3 BPA: Is It a Prototypical Endocrine Disruptor? The estrogenic activity of any ED compound is usually characterized by initial in vitro receptor binding and transactivation assays in an estrogenresponsive cell line. BPA exhibits relatively weak binding affinity for ER and ER , and similar results were observed in transactivation assays, suggesting an activity which was at least 1000-fold lower than observed for E2 (Kuiper et al. 1998). Moreover, in a relative proliferation potency assay in MCF-7 breast cancer cells (E-screen), BPA was 106 times less potent than E2 in some MCF-7 cells (Soto et al. 1995). Studies in this laboratory investigated the estrogenic activity of BPA in the ‘gold standard’ rodent uterotrophic assay in 21-day-old female rats, and the results showed that BPA at doses as high as 150 mg kg1 d1 (X3) did not increase uterine wet weight (Gould et al. 1998). However, like E2, BPA increased uterine progesterone receptor (PR) and

613

peroxidase activity, which suggested that BPA selectively activated uterine estrogenic responses. It was also observed that BPA in combination with E2 inhibited E2-induced PR binding and peroxidase activity but did not affect increases in uterine wet weight. These results suggested that, in the ‘classical’ estrogenic assays, BPA was both a weak ER agonist and a partial ER antagonist, which is consistent with an SERM (Gould et al. 1998). However, the reason for the conflicting opinions regarding adverse human health impacts of BPA are due to data from other studies, particularly those which involve in utero or early postnatal exposures to BPA (Vom Saal et al. 2007). Vom Saal and coworkers reported that in utero exposure to low doses of BPA (2 and 20 mg kg1 d1) induced an increase in prostate weight in 6-month-old adult males, whereas similar responses were not observed for octylphenol, and this was related to differences in relative binding affinity in the serum modified access assay. Similar low dose effects were observed for DES and EE, and other studies also reported BPA-induced responses at low doses (Nagel et al. 1997; Vom Saal et al. 1997). In contrast, other reports did not confirm results of these low dose studies (Ashby et al. 1999; Cagen et al. 1999) and, hence, the heated controversies regarding the potential adverse human health effects of BPA, in which accusations of bias have been raised by scientists on both sides of the issue.

2.31.4 Xenoestrogens and Phytoestrogens as SERMs: Implications for Risk Assessment 2.31.4.1 Multiple Pathways for ER Activation Steroid hormones induce their responses through interactions with their corresponding receptors which in turn interact with their cognate response elements in target gene promoters. Each hormone binds its own receptor with high specificity and has minimal affinity for other steroid hormone receptors (Nilsson and Gustafsson 2002; O’Malley 2005). However, the simple mechanism of action envisioned for ligand-activated nuclear receptors (including steroid hormone receptors) and this is exemplified by the high complex mechanisms of gene activation and repression induced by E2. For example, E2 can activate genes containing not only a consensus palindromic estrogen responsive element (ERE, GGTCA(N)3TGACC) but also promoters

614 Alterations in Cell Signaling

E2 ER Membrane

E2

Differe

nt kin

ER

ase pa

thway

s

ER

RE

ER

Sp1

ER

?

ER

GC

ERE

ERE1/2

1

148 AF-1

hERβ N

A/B

D

251

185

500 530 AF-2

53%

355

AF-1

DBD

Hinge

A/B

C

D

Nucleus

C

F

………

30%

Sp1

E/LBD

………

96%

ER ER

GC

214 304 Hinge DBD

………

…

………

…… hERα N

ER Jun

AP1

C

30%

1

ER Fos

ER

549

595

AF-2

E/LBD

F

C

Figure 3 Multiple mechanisms of ER-dependent activation of genes (top) and domain structures of ER and ER (bottom).

containing a variety of nonconsensus ERE or even ERE half-sites (Hyder et al. 1995). Moreover, several other mechanisms of ER-mediated gene expression have been identified and these include: (1) DNAbound ER interacting with other DNA-bound transcription factors such as specificity proteins (Sp); (2) DNA-independent pathways in which ER interacts with other DNA-bound transcription factors such as Sp proteins, AP-1, GATA, and NFB; and (3) extranuclear activation of ER (membrane and/or cytosolic) resulting in enhanced kinase signaling (Figure 3) (Blobel and Orkin 1996; Paech et al. 1997; Webb et al. 1999; Kalaitzidis and Gilmore 2005; Safe 2001; Watson et al. 2002). Thus, the simple mechanism of ER signaling has been greatly expanded to include multiple nuclear and extranuclear pathways. 2.31.4.2 Factors that Influence ER Activation The structures of ER and ER are modular like other members of the nuclear receptor superfamily and the various domains of these receptors

significantly increase the complexity of liganddependent ER signaling (Kuiper et al. 1997; Nilsson et al. 2001; Tora et al. 1989; Barkhem et al. 1998; Mosselman et al. 1996; Pettersson et al. 1997). Transactivation of ER or ER is dependent on N-terminal activation function-1 (AF-1) and C-terminal AF-2 domains, a hinge domain (D), and a DNA-binding domain (C) which contains two zinc finger motifs. The ligand-binding pocket for both ERs is within the C-terminal domain (E) and there is a sixth domain (F) at the C-terminal of ER and ER (Figure 3). Each of these domains has multiple functions which are required for ligand-dependent activation of ER. X-ray crystallographic analysis of the LBD of ER bound to various estrogenic compounds, such as E2, DES, the antiestrogens raloxifene and tamoxifen, and the pure antiestrogen ICI 164,384, an analog of fulvestrant (ICI 182,780), have been reported (Brzozowski et al. 1997; Pike et al. 1999, 2001; Shiau et al. 1998). The results show that differences between ER bound to agonists or antagonists is reflected in the structure of helix 12 within AF-2. When the ER LBD is complexed with E2, helix 12

Estrogenic Endocrine Disruptors: Molecular Characteristics and Human Impacts

is repositioned over the ligand-binding pocket and subsequently interacts with LXXLL motifs in various coactivators. Binding of ER antagonists such as raloxifene or tamoxifen to the ER-LBD results in displacement of helix 12 from its agonist positioning and this disrupts formation of the coactivator interaction surface (Pike et al. 1999; Shiau et al. 1998). The positioning of helix 12 is sensitive to ER ligand structure and, in part, dictates the extent of ERdependent activation (Brzozowski et al. 1997; Pike et al. 2001). Thus, ligand structure and its effect on conformational changes in ER will influence recruitment of coactivators, corepressors, and other coregulatory factors which are required for enhanced or repressed expression of a target gene. Moreover, the tissue-specific expression of coactivators, corepressors, and other cofactors will influence the activity of an ER ligand. In addition, promoter context and chromatin state which are dependent on histone methylation or acetylation and promoter methylation will also influence ligand-dependent activation of a specific gene. Due to all of these variables, ligands that bind ER can exhibit cell context-dependent ER agonist or antagonist activities and are therefore SERMs (Katzenellenbogen et al. 1996). For example, tamoxifen, a drug of choice for treating early stage ER-positive breast cancer, is an antiestrogen in mammary tumors but exhibits ER agonist activity in the bone, vascular system, and uterus (Brzozowski et al. 1997; Jordan 2003a,b; Katzenellenbogen et al. 1996; MacGregor and Jordan 1998; Pike et al. 1999, 2001; Shiau et al. 1998; Tora et al. 1989). There are also tissue- and speciesspecific ER agonist or antagonist activities of tamoxifen in animal models where this compound is an ER antagonist in chicks, a partial ER agonist/antagonist in rats, and an ER agonist in a number of short-term assays in mice (MacGregor and Jordan 1998). Studies in this laboratory have also demonstrated the importance of cell context on hormone-dependent gene activation or repression (Higgins et al. 2006, 2008). In ER-positive ZR-75 breast cancer cells, E2 induces vascular endothelial growth factor receptor 2 (VEGFR2) expression, whereas in ER-positive MCF-7 cells, E2 represses VEGFR2 expression. 2.31.4.3 SERMs Differentially Activate Wild-type and Variant ER The recognition that ligands that bind receptors such as ER exhibit tissue-specific agonist/antagonist activities leads to development of SERMs and other

615

selective receptor modulators for treating various diseases where the receptor can be a drug target (Jordan 2003a,b; Krishnan et al. 2000). Thus, if structurally diverse steroidal and nonsteroidal ER ligands such as tamoxifen, raloxifene, and fulvestrant are SERMs for treating breast cancer and hormonedeficient bone diseases, is it possible that structurally diverse xenoestrogens and phytoestrogens are also SERMs? If this is true, then estrogenic activity of phytoestrogen and xenoestrogen mixtures are not necessarily additive, and a prediction of the tissuespecific ER agonist or antagonist activity of an individual compound can only be obtained by actually measuring this response in an appropriate model. The structurally diverse SERMs, E2, tamoxifen, raloxifene, and ICI 164,384, have overlapping but also distinct in vivo biologies, and these compounds were used to develop an in vitro bioassay that would distinguish one from another (McDonnell et al. 1995; Tzukerman et al. 1994). This bioassay utilizes various domains of ER which exhibit both separable and overlapping functions that govern their interactions with other coregulatory proteins and promoter DNA. AF-1 and AF-2 are located in the A/B and E domains, respectively, and are particularly sensitive to the structure-dependent activities of SERMs. Activation of wild-type and mutant forms of ER has been extensively used in this laboratory to investigate the potential SERM-like activity of synthetic, xenoestrogens and, to a limited extent, phytoestrogens (Figure 4) (Gaido et al. 1999; Gaido et al. 2000; Gould et al. 1998, 2008; Higgins et al. 2006; Krishnan et al. 2000; Li et al. 2006; Tzukerman et al. 1994; Wu et al. 2008; Wu and Safe 2007; Yoon et al. 2000, 2001). Initial studies focused on activation of luciferase activity in different cell lines transfected with a construct (pERE3) containing three tandem EREs or an estrogen-responsive human complement C3 promoter (pC3-luc). The results were compared to those obtained for E2 or HO-TAM, the active form of the antiestrogen tamoxifen. A comparison of the structure-dependent activation of wild-type ER, ERAF1, or ER-AF2 in HepG2 liver cancer cell is summarized in Table 1. Maximal inducing concentrations of all compounds were used and, with few exceptions, a similar pattern of induction was observed in HepG2 and other cancer cell lines, although there were differences in percentage of maximal induction responses. In contrast, structureand cell context-dependent differences were observed for this same set of compounds in cells transfected with ER-AF1 which express the

616 Alterations in Cell Signaling

Figure 4 Structurally diverse estrogenic compounds used to activate wild-type and variant ER and ER/Sp ((_____); compounds in Figures 1 and 2 also included in this study).

Table 1 Structure-dependent induction of luciferase activity in HepG2 cells transfected with pERE3 and ER, ER-AF1, or ER-AF2 ER-AF1

ER Compounds 17 -Estradiol 4-Hydroxytamoxifen NP OP HO-PCB-CI4 HO-PCB-CI3 HPTE BPA Kepone Resveratrol

A/B

C D

E

þþþ þ þ þ þþ þþ þþþ þþ NI NI

F

A/B C D

þþþ NI þþ þ þ þ þ þ þ NI

ER-AF2 E

F

C

D

E

F

þþþ NI þþ NI þþ þþþ þþþ þþ þ NI

þþþ (100% response); þþ (40% of maximal response); þ (40% of maximal response); Nl (no significant induction).

full-length human ER with D538N, E542Q, and D545N mutations in helix-12 that inhibit ligandinduced coactivator interactions. For example, nonylphenol-induced transactivation in HepG2 cells but was inactive in U2 cells. In HepG2 cells, NP and OP activated luciferase activity in cells expressing ER and ER-AF1, whereas NP but not OP was active in HepG2 cells transfected with ER-AF2. Although there were both similarities and differences observed for activation of luciferase activity in cells transfected with wild-type or variant ER in HepG2 cells (Figure 5), it was clear that comparison of the results

over three cell lines (HepG2, U2, and MDA-MB231) showed that each compound exhibited a unique induction profile. 2.31.4.4 SERMs Differentially Activate Wild-type and Variant ER/Sp Studies in this laboratory have previously shown that in breast cancer cells, many E2-responsive genes are regulated by ER/Sp-dependent activation of GCrich sequences where Sp proteins but not ER bind promoter DNA (Porter et al. 1997; Safe and Kim 2004;

Estrogenic Endocrine Disruptors: Molecular Characteristics and Human Impacts

617

Table 2 Structure-dependent induction of luciferase activity in MDA-MB-231 cells transfected with pSp13 and ER, ER-
ZF1, or ER(1-537) ER Compounds

Concentration

17 -Estradiol Diethylstilbestrol 4-Hydroxytamoxifen ICI 182780 HPTE BPA OP NP HO-PCB-CI4 Resveratrol Endosulfan Kepone

10 nmol l1 10 nmol l1 1 mmol l1 1 mmol l1 25 mmol l1 75 mmol l1 25 mmol l1 10 mmol l1 25 mmol l1 75 mmol l1 25 mmol l1 10 mmol l1

A/B C D

þþþ þ þþ þ þþ þþ þþ þ þþ þ þ þ

ER-
ZF1 E

F

A/B

þ þ Nl Nl þþ þþ þ þ Nl þ þ þþþ

C D

ER(1-537) E

F

A/B C D E

Nl Nl þ þ þþ þþ þ þ þ þþ þþ þþþ

þþþ (100% response); þþ (40% of maximal response); þ (40% of maximal response); Nl (no significant induction).

Saville et al. 2000; Kim et al. 2003; Safe 2001). E2, DES, and the antiestrogens ICI 182,780 and HO-TAM activate wild-type ER/Sp in breast cancer cells transfected with pSp13, a construct containing 3 tandem GC-rich sites linked to the luciferase gene. However, E2 and DES but not the antiestrogens HO-TAM and ICI 182,780 activate transcription in breast cancer cells transfected with variant forms of ER containing deletions of the DBD, zinc finger 1 (ER
ZF1), or zinc finger 2 (ER
ZF2). Results in Table 2 show that in MDA-MB-231 cells transfected with pSp13 and ER, all xenoestrogens and resveratrol induced transactivation; however, in cells transfected with ER
ZF1, HO-TAM, ICI 182,780, and HO-PCB-Cl4 were inactive but kepone was the most potent inducer. However, in cells transfected with pSp3 and ER(1-537), a construct that lacks part of helix-12 in the LBD, E2, and DES do not induce transactivation, whereas the xenoestrogens, resveratrol, and the antiestrogens HO-TAM and ICI 182,780 all induce luciferase activity and kepone is the most active ligand. Thus, depending on the ER construct, the xenoestrogens resemble either DES/E2 or the antiestrogens in their activation of ER/Sp. We have also used RNA interference to knockdown Sp1, Sp3, and Sp4 to determine if there is a structuredependent activation of ER/Sp1, ER/Sp3, or ER/Sp4. For E2, DES, BPA, NP, endosulfan, and HPTE, Sp1 knockdown totally abrogates gene activation, Sp4 knockdown also decreases responsiveness,

whereas Sp3 knockdown had no effect. Thus, these compounds preferentially activate ER/Sp > ER/ Sp4 >> ER/Sp3. In contrast, transactivation induced by ICI 182,780 and kepone was decreased after transfection of small inhibitory RNAs for Sp1, Sp3, and Sp4; however, fold-inducibility was variable due to changes in basal activity. Thus, structure-dependent activation of ER/Sp is highly variable and depends on expression of wild-type or variant ER and Sp proteins and ongoing studies also show differences among different gene promoters. It is also clear from other studies that xenoestrogens and phytoestrogens are SERMs and therefore their tissue-selective ER agonist or antagonist activities are not readily predictable from simple receptor binding or transactivation assays. This could explain some of the unexpectedly high potencies observed for BPA in some assays. In addition, many estrogenic EDs also activate other endocrine-dependent and endocrine-independent responses that may contribute to the effects of these compounds. Despite the evidence showing that estrogenic ED compounds are SERMs, this does not necessarily mean that these compounds are adversely impacting human health. Moreover, there has always been some skepticism (Ahlborg et al. 1995; Safe 1995; Safe et al. 1997) that the hypothesized global increase in male reproductive tract problems or increased breast cancer incidence are related to exposure to ED compounds and this is discussed below.

618 Alterations in Cell Signaling

2.31.5 EDs and their Effects on Male Reproductive Tract Problems Sharpe and Skakkebaek initially hypothesized that diseases of the male reproductive tract pertaining to sperm counts and abnormalities, fertility, hypospadias, cryptorchidism, and testicular cancer were associated with in utero or early postnatal exposures to estrogens and/or antiandrogens (Sharpe and Skakkebaek 1993). This was later referred to as a ‘‘testicular dysgenesis syndrome (TDS): an increasingly common developmental disorder with environmental aspects’’ (Skakkebaek et al. 2001). Results of the initial meta-analysis of sperm counts from different locations concluded that ‘‘a significant decrease in mean sperm count from 113  106ml1 in 1940 to 66  106 ml1 in 1990 (p < 0.001)’’ (Carlsen et al. 1992) spurred a large number of sperm count studies worldwide to determine the validity of the hypothesis that ‘‘the increasing incidence of reproductive abnormalities in the human male may be related to increased estrogen exposure in utero’’ (Sharpe and Skakkebaek 1993). Results of sperm count studies in patients or selfselected volunteers and in some random samples have provided highly variable results, and this may be due, in part, to differences in measurement techniques and other factors. Nevertheless, in some studies, there are reports suggesting that sperm counts have decreased over limited periods of time (Safe 2002, 2004). For example, in a Parisian study of sperm donors, it was reported that ‘‘the mean concentration of sperm decreased by 2.1% per year, from 89  106 ml1 in 1973 to 60  106 ml1 in 1992 (p < 0.001)’’ and this study concluded: ‘‘During the past 20 years, there has been a decline in the concentration and motility of sperm and in the percentage of morphologically normal spermatozoa in fertile men that is independent of the age of the men’’ (Auger et al. 1995). In contrast, other studies examining sperm counts and their variation with time have demonstrated no apparent changes in sperm counts or semen quality. Moreover, a study in Toulouse, France, using a protocol comparable to that of the study in Paris showed that mean sperm count was 83.12  106 ml1 and that, from 1977 to 1992, ‘‘sperm concentration has not changed with time in the Toulouse area’’ (Bujan et al. 1996). A study by Marimuthu and coworkers examined sperm quality in over 1000 subjects from a national fertility clinic in New Delhi, India (Marimuthu et al. 2003). They

examined sperm collected over 11 years from 1990 to 2000 and showed that the average semen volume and sperm counts were 2.6  0.1  106 ml1 and 60.6  0.9  106 ml1. They concluded that ‘‘no significant decline in sperm counts was observed in any year during the entire study period’’ and ‘‘the present study has confirmed similar findings from other different countries that declining sperm counts in humans is not a global phenomena’’ (Safe 2002, 2004). In 1996, Fisch and coworkers ( 1996) investigated sperm quality of men from vasectomy clinics in California, New York, and Minnesota. Their data showed that sperm counts, sperm volume, and sperm motility remained unchanged from 1970 through 1994 in all three states. However, for the first time, they showed large differences in sperm counts in these locations. For example, sperm counts in New York, California, and Minnesota were 131.5, 72.7, and 100.8  106 ml1, respectively. These data demonstrated large demographic differences in sperm counts within the United States, and comparable observations have been reported in Canada and several European countries (Auger and Jouannet 1997; Gyllenborg et al. 1999; Jorgensen et al. 2002; Younglai et al. 1998). In Canada, sperm counts and quality were determined in 11 centers across the country and sperm counts ranged from 51 to 121  106 ml1 in 1984 and 48–137  106 ml1 in 1996 (Younglai et al. 1998). These results would suggest that environmental factors such as DDE and other trace OC contaminants and other xenoestrogens are unlikely to contribute to decreased sperm counts in Canada, the United States, or Europe since pollutant gradients within these regions are minimal (Ekbom et al. 1996; Kutz et al. 1991). Recent reports have examined twin pregnancies to probe genetic and environmental contributions to fertility and sperm quality. There is a correlation between decreased time to pregnancy (a measure of fertility) and twinning, and it was reported that fathers of twins had higher sperm counts and semen quality than a reference group (Asklund et al. 2007). In utero exposure to estrogens is much higher in twin pregnancies and also higher in dizygotic than monozygotic twins. Thus, if in utero exposure to estrogens is a critical element in the hypothesized decrease in male reproductive capacity, then twins with higher exposure to estrogens should exhibit decreased sperm counts. The authors reported that ‘‘higher prenatal concentrations of oestrogen are not related to reduced sperm counts in adulthood. In particular,

Estrogenic Endocrine Disruptors: Molecular Characteristics and Human Impacts

we did not find lower sperm counts in twin brothers: both the concentration and potency of oestrogens during pregnancy with twins are greater than for most environmental oestrogens’’ (Storgaard et al. 2002). Recent studies have presented an alternative explanation of the decrease in sperm quality in some regions such as the Copenhagen area in Denmark and this is related to body mass index. In a study of Danish young men examined for their fitness for military service, it was shown that men with a low (<20 kg m2) or high (>25 kg m2) body mass index had significant decreases in sperm concentrations and total sperm counts (Jensen et al. 2004). Similar results were obtained from a group of normal healthy men in the Atlanta, Georgia area (Kort et al. 2006). These results would suggest that the epidemic of obesity observed in many countries may be a critical contributor to decreased sperm quality in some regions and this is a condition that is preventable. Contributions of environmental factors to this condition are unclear.

2.31.6 Organochlorine EDs and Breast Cancer The initial concern regarding exposure to OC compounds and breast cancer was generated from two studies in Connecticut and New York that reported increased levels of DDE or PCBs in breast cancer patients versus controls (Falck et al. 1992; Wolff et al. 1993). The biological plausibility of the xenoestrogen-breast cancer hypothesis is questionable since PCBs/DDE are not mammary carcinogens in humans nor in laboratory animals, and some PCBs also exhibit antiestrogenic activity in rodent models (Mayes et al. 1998; Ramamoorthy et al. 1999; Scribner and Mottet 1981; Silinskas and Okey 1975). Moreover, although DDE exhibits antiandrogenic activity, the estrogenic activity of this compound is minimal. Subsequent studies in various cohorts throughout the world have shown no consistent correlations between serum or adipose tissue levels of PCB mixtures, PCB congeners and DDE, and an increased incidence of breast cancer in women (Gammon et al. 2002; Lopez-Cervantes et al. 2004; Safe and Papineni 2006). A recent report concluded that exposure of young women to p,p9-DDT during the peak use period of this insecticide in the United States (1950s and 1960s) was associated with increased risk for breast cancer (Cohn et al. 2007). The validity of this new hypothesis will have to be

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further tested; however, there are many countries in the Far East that have a long history of DDT use and have much lower rates of breast cancer. Moreover, the estrogenic activity of p,p9-DDT is minimal and the primary effects of this compound may be endocrine-independent.

2.31.7 Conclusions Although there is evidence that ED compounds may affect fish and wildlife populations, the evidence that these compounds impact human health is minimal, although some reports suggest a role for in utero exposure to phthalate plasticizers in mediating some antiandrogenic responses. Some scientists suggest that a host of human health problems may be mediated by BPA and these require further investigation since BPA and other xenoestrogens exhibit SERM-like activity. However, it should also be pointed out that some of the serious health changes in North American society such as precocious puberty are much less severe in highly developed countries such as Japan in which the Japanese population is exposed to a similar spectrum of ED compounds in the environment. Moreover, there is increasing evidence that diet and diet-induced obesity represent a serious health problem in some developed countries such as the United States. Therefore, the effects of ED compounds on humans will have to be evaluated and compared to the effects of a diet that enhances obesity and also a diet that contains relative high levels of estrogenic phytochemicals that are generally associated with health benefits.

Acknowledgments The financial support of the National Institutes of Health (ES04917) is appreciated.

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