Environmental toxicants and effects on female reproductive function

Abstract

One of the most toxic substances known to humans, 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD or dioxin), is also highly pervasive in the environment. It is created naturally in volcanic eruptions and forest fires, and anthropogenically in waste incineration, chlorination processes and certain plastics manufacture. From reports of large industrial and other accidents, or from experimental studies, dioxin exposure has been correlated in animal models and/or humans with chloracne of the skin, organ cancers, hepatotoxicity, gonadal and immune changes, pulmonary and other diseases such as diabetes, skewing of the sex ratio, and infertility. We have demonstrated that the aromatic hydrocarbon receptor (AHR) that binds dioxin in tissues is localized in zebrafish, rat and rhesus monkey (Macaca mulatta) ovaries and in rat and human luteinizing granulosa cells (GC) (among other tissues), that labeled dioxin is specifically localized to granulosa cells of the ovarian follicle as observed by autoradiography, and that incubations of GC or ovarian fragments with environmentally relevant concentrations (fM to nM) of dioxin inhibit estradiol secretion significantly. Our experiments show that in human, non-human primate, rat, trout, and zebrafish ovarian tissues, dioxin inhibits estrogen synthesis at some level of the steroid biosynthetic pathway, most likely by inhibiting transcription of mRNAs for or activity of side-chain cleavage (Cyp11a1 gene) and/or aromatase (Cyp19a1 gene) enzymes, or conceivably other steroidogenic enzymes/factors. Such an untoward effect on estrogen synthesis in females exposed to dioxin environmentally may predispose them to defects in aspects of their fertility.

INTRODUCTION

The premise that there exist man-made substances that accumulate in the environment and which are responsible for reproductive and developmental deficits observed in animals was brought to our attention vividly in the early 1960’s by Rachel Carson in her ground-breaking book, Silent Spring [1]. The field of reproductive toxicology was ultimately created as a result of this and other work, and has advanced rapidly in recent years especially because of the widening public interest in environmental pollutants that are endocrine disrupting or that otherwise modulate in some way the reproductive system of animals and humans. Particularly in the past 10 years, researchers from various disciplines have lent their expertise to reproductive toxicology, hastening the development of this area. Physiologists, endocrinologists, molecular biologists, biochemists and others that work in the field of reproduction have begun to evaluate xenobiotic compounds and their effects on the reproductive axis (hypothalamus, anterior pituitary, testis and ovary, uterus and male sex glands). This review will focus on effects of environmental pollutants on female reproductive function.

Reproductive hazards in the environment

Reproductive deficits induced by estrogenic substances in the environment constitute one such recent concern. We have incurred large estrogenic exposures due to diethylstilbestrol (DES), plant or phytoestrogens, and other man-made estrogen-like chemicals [2]. There have been numerous reports that revealed high concentrations of contraceptive or menopausal estrogens in wastewater effluent in English rivers; and these may be responsible for alterations in fish sex ratios [3, 4]. In addition, human populations carry PCBs due to diets high in fish heavily laden with these compounds [5]. Certain PCBs, although only loosely linked to diminished fertility rates in humans, are quite strongly documented with regard to reproductive impairment in birds, fish and mammals [6]. The pesticide DDT, which is still used in developing countries, has been shown to exert deleterious effects on wildlife populations, and might be linked to breast and other cancers in women exposed to this pesticide as long as 20 years ago [7–9].

Some Great Lakes’ fish show reduced gonadal weights, sex steroid hormone concentrations, and survivability [10], due potentially to the wood-derived phytoestrogen β-sitosterol, dioxins, PCBs, etc.[11]. Whether these effects are the result of endocrine disruption is not yet clearly established, as lake trout contaminated with PCBs were fed to rats, resulting in few endocrine changes except for subtle alterations in aromatase activity (which converts androgens to estrogens) [12].

Regarding other reproductive toxicants, Baldridge showed reduced numbers of small preantral follicles and of mid- and large-sized antral follicles with the PCB mixture Aroclor 1016; and that this effect was negated with thyroxine supplementation only in the preantral follicles, and “reversed” signs of atresia/apoptosis in the smallest antral follicles [13]. Similar effects were observed with ammonium perchlorate (found in explosives and rocketry materials) [14]. Dioxins and β-naphthoflavone (BNF, both AHR agonists) reduced vitellogenin synthesis in fish primary hepatocyte cultures [15] and demonstrated a negative correlation between cytochrome P450 1A1 (Cyp1a1)-inducing capability (a prominent biomarker of dioxin/AHR agonist action) and anti-estrogenicity. This fits with our results where presumed greater Cyp1a1 activity with increasing BNF dosage administered to rainbow trout resulted in reduced circulating E₂ concentrations in a dose-dependent manner at 48 h of exposure [16], but without affecting genes involved in immune regulation (a typical sequela of AHR agonism) [17].

It is tempting to overstate the contention that all environmental compounds exert a negative effect on physiologic systems in animals. However, the effects observed are complex and dependent upon the compound evaluated, whether it is in a binary or other mixture, whether its concentration is environmentally relevant (i.e., similar to that found naturally), the route of exposure/administration, the animal model evaluated, the experimental paradigm used (e.g., in vitro, in vivo), etc.

Although our laboratories have investigated effects of endogenous and xenobiotic estrogens for almost 25 years, we have since about 1993 evaluated other environmental molecules and toxicants that appear to alter the ER-signaling pathway (e.g., dioxin) and thereby modulate ovarian function and female reproduction generally. Herewith, we have attempted to discern a role(s) for alterations in ER signal transduction or in estrogen biosynthesis in the untoward effects observed on fertility and more subtle indices of reproductive capacity. A drop in circulating estrogen or in its action may be responsible for the diminution in fertility that is characteristic of captive and wild animals exposed to certain toxicants such as TCDD [18].

TCDD has been referred to as the most toxic substance known to man and is produced as a by- product of waste incineration (personal now more so than municipal), herbicide overuse, paper chlorination, polyvinylchloride plastics production, etc., and we are exposed to dioxin-containing animal products via our daily diet [18, 19]. Aside from acting as a teratogen, embryotoxin, carcinogen (the Report on Carcinogens [20] listed TCDD as “known to be a human carcinogen”), tumor progressor, and innate and adaptive immune function suppressor, TCDD is a known endocrine disrupter in laboratory species, wildlife, and from the data from several accidental exposures, in humans [18]. We have demonstrated attenuated secretion of ovarian estrogen, reduced numbers of large preovulatory ovarian follicles and altered expression of ovarian mRNAs for ER and AHR in female peri-pubertal rat pups after in-vivo administration of TCDD to pregnant dams [21, 22]. These results on reduced E₂ secretion and reduced antral follicle numbers are now corroborated by several recent reports [23–25]. We have more recently demonstrated that TCDD reduces Cyp11a1 and Cyp19a1 mRNA copy numbers in rat after equine chorionic gonadotropin (eCG)-induced ovarian superstimulation using competitive RT-PCR [26]; although another group has suggested little effect of TCDD on human GC CYP19a1 [24, 27]. Regardless, the available evidence (especially in vitro) strongly supports the ovary as a major target of TCDD action [18, 28, 29]. The putative mechanism for this action is described below.

Mechanism for TCDD action

Most of the biologic effects of TCDD are mediated upon binding to a high-affinity cytoplasmic receptor, the AHR (refer to Figure 1 below) [30]. Binding of ligand to cytosolic AHR (roughly 800 amino acids long, a ligand-inducible transcription factor and member of the PER-ARNT-SIM [PAS] transcription factor family) induces allosteric modifications in the AHR, and the receptor-ligand complex undergoes conformational changes including dissociation of two heat-shock proteins (HSP-90s) and an AHR-interacting protein (AIP), and/or other protein(s), depending upon cellular paradigm. Ligand-bound AHR is then translocated to the nucleus where it further complexes with the resident ARNT (or hypoxia-inducible factor [HIF] 1β) protein and, in certain cell types, the retinoblastoma protein (RbP, [31]). This heterodimer forms a transcriptional complex at the AHRE (TNGCGTG DNA motifs, concensus binding sequence underlined) in the regulatory region of several target genes (e.g., cytochromes P450 Cyp1a1, Cyp1a2, Cyp1b1) and some phase II detoxifying enzymes, inducing gene transcription. White and Gasiewich [32] reported that the ER structural gene (Esr) contains several AHREs, and therefore may be a site of interaction (“crosstalk”) between signaling pathways; e.g., Ohtake et al. have shown that unliganded ER can be affected by liganded AHR agonistically [33]. To complicate matters further, it appears that estrogen also modulates the cellular response to TCDD. In MCF-7 cells (human breast carcinoma), TCDD inhibited estrogen-induced cathepsin D (CTSD) expression whereas TCDD induction of CYP1A1 was inhibited by estrogen [34]. Tian et al. [35] reported that TCDD down-regulates Esr1 mRNA in both the liver and ovary of female CD-1 mice with the greatest effect in the ovary. Therefore, there does appear to exist reciprocal receptor crosstalk. This cross-talk may predispose effects on fertility.

Figure 1. Simplified schematic of AHR activation by TCDD.

TCDD binds the AHR in the cytoplasm, which displaces heat shock proteins and an AHR-interacting protein. TCDD-AHR translocates to the nucleus and is complexed with ARNT. This heterotrimeric complex binds AHREs or iAHREs and interacts with co-activators or co- represssors at the enhancer regions of DNA to achieve biologic effect (transcriptional activation or inhibition). Interactions with the ER may occur in multiple regions of the cell and this may represent a potential site of cross-talk between AHR- and ER-signaling pathways.

Dioxin and female fertility

We have chosen to focus our energies therefore on the reproductive effects of dioxins. Recent reports have concentrated on TCDD’s reproductive and antiestrogenic effects. In rats, TCDD has been shown to interfere with the maintenance of pregnancy, fetal growth, and development, and fecundity and fertility parameters [36–38]. In primates, TCDD exposure resulted in significant hormonal alterations, failure of female rhesus monkeys to conceive, and was correlated with increased severity of endometriosis [39, 40]. Umbreit and Gallo suggested that TCDD’s effects on reproductive functions are due to its effects on the actions of estrogens [41]. The antiestrogenic effects of TCDD were observed as a reduction of ER levels in immature mice [42]; TCDD reduced basal and E₂-stimulated ER in rats [43–45]. In MCF-7 cells, TCDD has also been noted to reduce ER levels [46]. Others, however, have shown that the effects of TCDD on ER are both species and tissue specific, and dependent upon the age and hormonal status of the animal [47, 48].

Petroff et al. [49] using an immature rat model showed in vivo that E₂ administered locally antagonized the TCDD effect of modulating ovulation. And it now appears that TCDD in vivo reduces the number of large preantral/antral ovarian follicles, and that this may be the reason for the reduced peripheral E₂ concentrations [21, 22]. (Interestingly, this effect is also shown with AHR null or “knockout” mice; i.e., a 50% drop in the number of pre-antral/antral follicles in the ovaries of deficient mice at 53 days of age [23]). Genetic inactivation of AHR may emulate administration of TCDD, as we observe a down-regulation of Ahr mRNA with TCDD administration [50]. And although 2 μg/kg TCDD was shown to increase the number of days to first litter in Siberian hamsters [51], no long-term fertility defect was observed in this model. Finally, nM TCDD significantly reduced P₄ accumulation into culture medium by porcine luteal cells [52]. Recent work showed an overall reduction in ovarian cAMP accumulation with TCDD in vivo, and a reduced binding of FSH and hCG to rat GC [25]. This implies a local inhibition by TCDD in reducing ovarian response, which is further supported by the attenuated follicular development and diminished pre-antral/antral follicle numbers noted above and observed by ourselves and others [21, 22, 25].

In studies that followed up those of Peterson and others [53], we have demonstrated a reduction in circulating estrogen by about 50% in female rat pups after maternal oral administration of 1 μg TCDD/kg body mass [21]. This is accompanied by an increase in pup ovarian and uterine Esr1 mRNA, some diminution in mRNA for β-subunit of FSH (Fshb), and changes in ER and presumably in estrogen-signaling pathway function observed in hypothalamus and pituitary. Dasmahapatra et al. showed increased steady-state Esr2 mRNA with TCDD in rat GC, and some augmented metabolism (hydroxylation) was expected because of increases observed in Cyp1b1 [54]. We have shown that the presence and expression of Ahr and Arnt mRNAs correlated with changes in the rat estrous cycle [50]; and have localized them to monkey ovary and hLGC [55]. An index of AHR activation is increased Cyp1a1 and Cyp1b1 messages and proteins [54, 56]. We have also very recently demonstrated topographic localization of presumed dioxin effect by using autoradiography of rhesus monkey ovaries. The ovaries bound radiolabeled dioxin in the granulosa cells and oocytes of their antral follicles, and this specific binding could be blocked by the addition of 200-fold excess of the AHR antagonist, α-naphthoflavone [57]. TCDD also appeared to inhibit mid-sized ovarian follicle growth, but did not accomplish this by enhancing GC apoptosis in the rat [22]. And with regard to effects on embryonic development, TCDD did not retard the development of early mouse morulae to blastocysts, and did not alter degree of apoptosis of these early pre-implantation embryos [58]. Other types of in-vitro studies allowed for more profound experimental manipulation.

Similar to the studies with incubated rat GC, we used cultures of GC aspirated from ovarian follicles in women participating in in-vitro fertilization protocols [59, 60], and attempted to delve deeper into the mechanism(s) whereby ovarian estrogen secretion was reduced. Heimler [59] showed a dramatic reduction in GC secretion of E₂ in vitro by 8 h of culture with as little as 3.1 pM TCDD. This reduction was prevented with the inclusion of an aromatizable substrate, A₄, at 10⁻⁷ M, suggesting that aromatase activity per se was not affected, but rather provision of the androgen precursor (cf. Figure 2). More recently, we have demonstrated that fM concentrations of TCDD (the lowest yet reported) substantially reduced E₂ output (Baldridge et al., unpublished observations; Figure 3) by hLGC and increased inhibin A secretion [60]; this may be the reason why we and others have observed a diminution in Fshb mRNA [61] or in its receptor (indicative of FSH responsiveness) [29], as inhibins feed back negatively on pituitary FSH secretion. Moran et al. showed in 2000 and 2003 that indeed in hLGC cultured with hCG long term, CYP17A1 activity (which provisions androgens) was compromised, and that this is a locus of major effect in the steroidogenic pathway in their system [24, 27]. However, it certainly appears that the effect is model dependent, as we have shown a diminution in Cyp19a1 mRNA and activity and Cyp11a1 mRNA in rat [26]. Therefore, it is clear that there is species dependence in dioxin effects, although the diminution in ovarian estrogen secretion is a common thread. We have now shown repeatedly that estrogen is attenuated in several animal systems: human LGC [59], rat [21, 26], trout [16] and zebrafish [62]; and we showed recently that cultures of ovarian fragments from rhesus monkey ovary manifested reduced secretion of E₂ after 48 h of culture [63] (Figure 4). We are currently further evaluating in rat the promotor regions within the genes of the two cytochromes involved in order to understand the potential regulation of steroidogenesis by TCDD [64, 65]. To bring our in-vitro studies full circle, we have also been developing an in-vivo zebrafish model.

Figure 2. Brief schematic of the steroidogenic pathway to estrogens in ovarian GC.

Cholesterol is transported to the inner mitochondrial membrane of ovarian follicle thecal cells by Star through the peripheral benzodiazepine receptor channel in the outer mitochondrial membrane, and cleaved to P₅. Thecal smooth endoplasmic reticulum is then the site for Hsd3b isozyme conversion of P₅ to P₄ and DHEA to A₄; and for the conversion of P₅ to DHEA under Cyp17a. A₄ ultimately diffuses across the basement membrane to layers of mural granulosa cells where it is aromatized by Cyp19a1 to estrone; and interconverted with E₂ via Hsd17b isozymes, as occurs between A₄ and testosterone.

Figure 3. Effects of TCDD on E₂ secretion by human LGC.

TCDD at fM, pM and nM concentrations caused a significant decrease in E₂ secretion at 8, 12 and 24 h in all treatment concentrations versus control. Data are presented as mean ± S.E.M. Letters (a, b) denote significance among treatment groups, based on Tukey’s post-hoc test (P < 0.05). Treatments sharing the same letter did not differ significantly. N = 8 for each treatment group.

Figure 4. Comparison of average E₂ release noted in monkey ovarian fragments cultured with either control medium (CTL) or medium containing pM TCDD (TCDD).

TCDD exposure at this environmentally relevant concentration markedly suppressed E₂ release at 24 (*, p < 0.003) and 48 (*, p < 0.02) h. Although marked suppression in E ₂ release was also suggested at 72 and 96 h, differences between the CTL and TCDD cultures were not found to be statistically significant (p = 0.07 and p = 0.20, respectively). Rather, variability of response at these latter times increased dramatically. Data represent means ± SE. “n” = 3 at 24, 48, and 72 h; “n” = 2 at 96 h.

We instituted the use of an in-vivo model, that of the genetically well-characterized zebrafish. A major advantage here of this relatively simple in-vivo model is the ability to apply sublethal exposure over the lifetime of the animal. We orally exposed adult female zebrafish to 0.08 to 2.16 ng TCDD/fish/day and observed accumulation of TCDD in a dose- and time-dependent manner in all tissues investigated but brain (1.1 – 36 ng/g) [66]. The ratio of ovarian to body mass (ovosomatic index) was reduced with TCDD concentrations of 0.6 ng/g fish or above; and 10% of females showed ovarian necrosis with 3 ng/g. The lowest observed adverse effects level (LOAEL) for female zebrafish was 0.6 ng/g fish, and the LOAEL for effects in their offspring was 1.1 ng/g fish. The maternal transfer of the TCDD induced larval toxicity similar to blue sac syndrome, which is a disorder of the nutrient-rich yolk sac of developing fish. Although zebrafish are overall less sensitive to dioxin than are other fish, their sensitivity may actually be more similar to that of mouse, rat, and human [66].

Egg production in zebrafish was decreased by over 50% following exposure to 40 and 100 ng/g in the diet for 5 days, and spawning success (females able to reproduce) was decreased by up to 96% [62]. Serum E₂ was also reduced more than two-fold, which is very similar to that observed for other animal/human/in-vitro model systems. Following 15 days of dietary exposure to 0.08 to 0.8 ng TCDD/fish/day, we also evaluated the expression of genes important in estrogen synthesis and signaling. TCDD exposure decreased significantly mRNA abundance for FSH receptor, LH receptor, Esr1, Esr2a and Esr2b, and greatly diminished Star (which translocates cholesterol to the mitochondrion), Cyp19a1, activins and epidermal growth factor receptor. There was a reduction in the total numbers of follicles after 20 days at 100 ng/g TCDD and a dose-dependent increase in atretic follicles. Fewer large or vitellogenic follicles were seen with treatment and there were significantly fewer secondary growth follicles; vitellogenic follicles only remained in the lowest treatment group. Although large deficits were incurred in the ovary with TCDD, whether the defect is entirely at this level is controversial.

We expect that to a great extent the steroidogenic defect is at the ovarian level, but our and others’ studies do not preclude effects at additional loci within the reproductive axis [49, 61, 67]. Therefore, effects could be exerted at multiple levels. To this end we have shown subtle effects on Fshb mRNA in pituitary but not on dimeric LH or FSH in circulating blood [61]. We have to date not observed effects of TCDD in vitro on pulsatile release of gonadotropin-releasing hormone (GnRH) from hypothalamic explants in rat, nor on LH or FSH secretion from hemi-pituitary cultures [68, 69]. Others have shown that TCDD alters pituitary responsiveness to GnRH from the hypothalamus [70]; and may modulate action of Star [71], and/or alter Cyp11a1 enzyme action [72]. Mutoh et al. [73] recently supported this by demonstrating that dioxin inhibited cholesterol translocation and steroidogenesis in rat by reducing synthesis of fetal brain LH and expression of testicular Star. In males this may be one mechanism of TCDD action at the steroidogenic pathway. Whether this holds for females also is unknown presently. Although several loci may be affected in subtle fashion by TCDD, it is obvious that ovarian steroidogenesis is greatly altered--we believe directly. A brief outline of steroidogeneis in the ovary follows.

Steroidogenic pathway

It is known that inner mitochondrial membrane of ovarian follicle thecal cells is the location of the Cyp11a1 and that cholesterol is transported there by Star through the peripheral benzodiazepine receptor channel in the outer mitochondrial membrane, and cholesterol is then cleaved to P₅ (refer to Figure. 2). TCDD inhibited rat testis steroidogenesis by diminishing mobilization of cholesterol to Cyp11a1, a clear indication that a TCDD effect on Star protein is implicated in this model [71, 73]. Thecal smooth endoplasmic reticulum (microsomal fraction) is then the site for Hsd3b conversion of to P₅ P₄, and dehydroepiandrosterone (DHEA) to A₄; and for the conversion of P₅ to DHEA under Cyp17a1. A₄ ultimately diffuses across the basement membrane (basal lamina) to layers of mural granulosa cells (membrana granulosa), where it is aromatized by Cyp19a1 to estrone and interconverted with E₂ via Hsd17b1 and Hsd17b2.

Hypothesis for TCDD action

Our overall hypothesis has been that TCDD exerts an anti-reproductive effect by binding AHR and affects ovarian steroidogenesis at two loci in the steroid biosynthetic pathway; i.e., at the level of Cyp11a1 and Cyp19a1, specifically. We have previously evaluated several of the mRNAs/enzymes involved in the steroidogenic pathway outlined below in Figure 2.

We have observed no effects of TCDD on Cyp17a1 or Hsd3b1 mRNA copy number in rat GC (refer to Fig. 2), but possess unique data that Cyp11a1 and Cyp19a1 mRNA copy numbers are drastically reduced in GC from rats that are either unstimulated (basal condition) or stimulated with exogenous gonadotropins (e.g., FSH) in the presence of environmentally relevant concentrations of TCDD [26]. This is supported by our very recent zebrafish data, where both Cyp11a1 and Cyp19a1 mRNAs were also reduced with TCDD in vivo as ascertained using real-time RT-PCR [64]. Our aims have been to show effects using low-dose TCDD, and to dissect these effects at the molecular level.

Potential transcriptional effects

It is known that AHRE sequences function as transcriptional enhancers at the Cyp1a1 gene [74, 75]; however, AHRE-like sequences are also implicated in gene suppression in genes such as pS2, Ctsd, and c-fos. In each case, the AHRE sequence contains the four invariant nucleotides (the CGTC “core”) required for AHR binding [5′-ACGTGNN(A/T)NNN(C/G)-3′], but contains one or more mismatches with the consensus sequence for functional activity (i.e., dioxin responsiveness, [5′-(T/G)NGCGTG (A/C)(G/C)A-3′]), and additional flanking sites (creating a 9-total-bp sequence) may play a role. Binding of the agonist-bound AHR to these “inhibitory” or “iAHREs” is thought to interfere with the binding of other positive transcriptional activating factors to nearby response elements (e.g., steroid receptor co-activator [SRC]-1 to transactivation function [AF] sites, [76]), thus resulting in down-regulation of gene expression. Previous work also shows that AHR cooperates with an orphan nuclear receptor, Ad4BP/SF-1, to activate Cyp11a1 gene transcription in ovarian GC [77], thereby providing more evidence of transcriptional effects of TCDD. We are currently evaluating in our zebrafish model the in-vivo effects of dioxin and searching steroidogenic enzyme promotor regions for these (i)AHRE’s. Binding of dioxin/AHR/ARNT to these (i)AHREs may constitute a dosage effect and result in gene activation or repression, but would not preclude other mechanism(s) of TCDD action on estrogen secretion.

The above findings are primarily descriptive and much research is still needed before we understand the molecular regulation of reproductive function by dioxin and other AHR agonists. We summarize our overall recent findings as follows.

SYNOPSIS

The ovaries of several species bind TCDD via the AHR in the GC primarily although certainly not exclusively. Ovarian E₂ secretion is diminished following TCDD (or other TCDD-like molecule) exposure in vivo (rat, trout and zebrafish) and in vitro (human and rat GC and monkey ovarian fragments). Biochemically, this deficit is manifested in the disruption of the steroidogenic pathway, resulting in attenuated secretion of estrogen particularly. In-vitro work suggests less of an effect at the pituitary or hypothalamus, but rather a direct, local, i.e, ovarian, effect. In several model systems, this impairment is at the level of transcription of several steroid-synthetic enzymes, Cyp11a1 and Cyp19a1, for example. While the biochemical effects can certainly differ with respect to dose, duration, animal, or model system, we believe that the resultant inhibition of estrogen synthesis is then responsible for the reduced fertility of animals or humans exposed to dioxins in the environment.

The significance of this recent work is the focus on very low levels of the “real-life” toxin TCDD to which animals and women are exposed during their reproductive years; encountering TCDD in their surrounding environment from food, air, water, chemicals, accidental exposures, etc. Evidence suggests that environmental pollutants such as dioxins can engender fertility-disrupting deficits in animals and humans. Further research is expected to continue to address the significant problem of potential reproductive hazards posed to women and animals by TCDD and other noxious environmental chemicals.

ABBREVIATIONS

A₄

androstenedione

AHR

aromatic hydrocarbon receptor

AHRE

AH-responsive element

ARNT

AHR nuclear translocator protein

Cyp

rat (lower case italicized

human

upper case italicized) cytochrome P450 gene, cDNA, and mRNA

Cyp11a1

P450 C_20,22-side-chain cleavage gene, cDNA, and mRNA

Cyp11a1 (non italicized)

enzyme protein

Cyp19a1

P450 aromatase gene, cDNA, and mRNA

Cyp19a1

enzyme protein

Cyp17a1

P450 C_17,20-lyase gene, cDNA, and mRNA

E₁

estrone

E₂

estradiol-17β

Esr1

Esr2a, Esr2b, ER, estrogen receptor (α, β1, β2) genes, protein

ERE

estrogen-responsive element

Fshb

follicle- stimulating hormone gene

GC

granulosa cells

h

hours

hCG

human chorionic gonadotropin

hLGC

human luteinizing granulosa cells

Hsd3b

Hsd3b, 3β-hydroxysteroid dehydrogenase

Hsd17b

Hsd17b, 17β-hydroxysteroid dehydrogenase

PCB

polychlorinated biphenyl

P₅

pregnenolone

RT-PCR

reverse transcriptase-polymerase chain reaction

Star

Star, steroid acute regulatory protein gene, protein

TCDD

tetrachlorodibenzo- p-dioxin

fM

pM, nM, femto-, pico-, nanoMolar