![Structural formulas of vinclozolin, vinclozolin metabolites M1 and M2, DHT, DDE, and hydroxyflutamide& / f i g. c : Fig. 3 Inhibitory effects of increasing concentrations of vinclozolin and metabolites M1 and M2 on AR-mediated gene transcription. Transcriptional activity was determined in CV1 cells by transiently transfecting human AR expression vector pCMVhAR and a mouse mammary tumor virus promoter-luciferase reporter vector. Cells were treated with 0.1 nM DHT with or without the indicated concentrations of competing ligands. Luciferase optical units are shown with standard error, and fold induction is indicated (bottom) relative to the activity determined in the absence of 0.1 nM DHT. p5 represents results when the cells were cotransfected with the luciferase reporter vector and the parent AR expression vector pCMV5, which lacks AR sequence. (Modified from [38])& / f i g. c :](https://www.researchgate.net/profile/Elizabeth-Wilson-12/publication/14111857/figure/fig2/AS:669308183146518@1536586914669/Structural-formulas-of-vinclozolin-vinclozolin-metabolites-M1-and-M2-DHT-DDE-and_Q320.jpg)
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J Mol Med (1997) 75:198–207 © Springer-Verlag 1997
&p.1:Abstract Industrial chemicals and environmental pol-
lutants can disrupt reproductive development in wildlife
and humans by mimicking or inhibiting the action of the
gonadal steroid hormones, estradiol and testosterone.
The toxicity of these so-called environmental endocrine
disruptors is especially insidious during sex differentia-
tion and development due to the crucial role of gonadal
steroid hormones in regulating these processes. This re-
view describes the mechanism of toxicity and clinical
implications of a new class of environmental chemicals
that inhibit androgen-mediated sex development. For
several of these chemicals, including the agricultural
fungicide vinclozolin and the ubiquitous and persistent
1,1,1-trichloro-2,2-bis (p-chlorophenyl)ethane metabo-
lite, 1,1-dichloro-2,2-bis(p-chlorophenyl) ethylene, the
molecular mechanism of action and the adverse devel-
opmental effects on male sex differentiation have been
elucidated and are used as examples. Environmental
chemicals with antiandrogenic activity offer profound
implications with regard to recent clinical observations
that suggest an increasing incidence of human male gen-
ital tract malformations, male infertility, and female
breast cancer. Finally, in light of increasing concern
over the potential endocrine disrupting effects of envi-
ronmental pollutants, an in vitro/in vivo investigational
strategy is presented which has proved useful in identi-
fying chemicals with antiandrogen activity and their
mechanism of action.
&kwd:Key words Antiandrogen · Endocrine disruptor ·
Sex differentiation · Androgen receptor ·
Hypospadias · Breast cancer · Vinclozolin ·
1,1-Dichloro-2,2-bis(p-chlorophenyl) ethylene
Abbreviations AR Androgen receptor ·
DDE 1,1-Dichloro-2,2-bis(p-chlorophenyl) ethylene ·
DDT 1,1,1-Trichloro-2,2-bis (p-chlorophenyl)ethane ·
DHT 5 -Dihydrotestosterone · ER Estrogen receptor ·
M1 2-[[(3, 5-Dichlorophenyl)-carbamoyl] oxy]-2-methyl-
3-butenoic acid · M2 3′,5′-Dichloro-2-hydroxy-
2-methylbut-3-enanlide · PR Progesterone receptor&bdy:
Introduction
Steroid hormone receptors control fundamental events in
embryonic development and sex differentiation through
their action as ligand-inducible transcription factors that
activate or repress transcription of target genes. The con-
sequences of disrupting these processes can be especially
profound during development due to the crucial role of
steroid hormones in controlling transient and irreversible
developmental processes. We are beginning to realize
that certain industrial pollutants and the widespread use
of pesticides and fungicides in the environment have the
potential to alter male sex development and reproductive
processes in wildlife and human populations by acting as
environmental antiandrogens [1]. In some cases laborato-
ry studies have confirmed abnormalities of reproductive
development observed in the field and have provided
mechanisms to explain the disruptive effects of these en-
vironmental chemicals. Included in this list are recent
studies with the ubiquitous environmental contaminant
1,1-dichloro-2,2-bis(p-chlorophenyl) ethylene (DDE), the
major stable metabolite of 1,1,1-trichloro-2,2-bis (p-
chlorophenyl)ethane (DDT), which exhibits antiandro-
genic activity [2]. Although environmental chemicals as
antiandrogens have been reported previously [3, 4], our
studies have focused attention on this new class of envi-
ronmental endocrine disruptors.
W.R. Kelce (
✉
)
Reproductive Toxicology Division, Endocrinology Branch,
National Health and Environmental Effects Research Laboratory,
United States Environmental Protection Agency,
Research Triangle Park, North Carolina 27711;
and Laboratories for Reproductive Biology
and the Department of Pediatrics, University of North Carolina,
Chapel Hill, NC 27599, USA
E.M. Wilson
Laboratories for Reproductive Biology
and the Departments of Pediatrics and of Biochemistry
and Biophysics, University of North Carolina, Chapel Hill,
NC 27599, USA&/fn-block:
REVIEW
&roles:William R. Kelce · Elizabeth M. Wilson
Environmental antiandrogens: developmental effects,
molecular mechanisms, and clinical implications
&misc:Received: 24 June 1996 / Accepted: 4 September 1996
199
Reports of the increasing incidence of developmental
reproductive tract abnormalities (e.g., hypospadias) in
the human male population and decreased adult sperm
counts in some parts of the world, together with the re-
cent identification of major pesticides that have antian-
drogenic activity, necessitates a closer look at the mole-
cular, functional, and clinical implications of antiandro-
gens in the environment. To this end we (a) review the
proposed mechanism(s) whereby androgen receptor
(AR) agonists and antagonists induce and/or inhibit
transcription of AR-dependent genes, (b) discuss the de-
velopmental effects and molecular mechanisms of sever-
al recently discovered environmental antiandrogenic
chemicals, and (c) discuss the clinical implications of
antiandrogenic chemicals in the environment. Where ap-
propriate, we relate what is known regarding the mole-
cular mechanisms of other steroid hormone receptors to
support generalizations related to AR. A better apprecia-
tion for the mechanisms involved in antiandrogen action
will be applicable to a number of clinical settings in-
cluding the increasing incidence of male genital tract
malformations, male infertility, and the increasing inci-
dence of tumorigenic effects of estrogens in women. Fi-
nally, as opinions vary regarding the best approach to
identify endocrine disruptors, i.e., primarily in vitro ver-
sus in vivo assays, we have compiled a comprehensive
investigational strategy to identify chemicals and/or me-
tabolites and the mechanism responsible for the pheno-
typic effects.
Mechanisms of androgen/antiandrogen action
Testosterone dissociates from carrier proteins in human
plasma and diffuses into cells where it binds intracellular
AR. In some target tissues such as prostate and epididy-
mis testosterone is reduced to the more potent androgen
5-dihydrotestosterone (DHT) by the microsomal enzyme
5-reductase (E.C. 1.3.99.5). The enhanced androgenic
activity of DHT results in part from its slower rate of dis-
sociation from AR [5]. High-affinity androgen binding
induces a conformational change in the AR that is requi-
site for stabilization against proteolytic degradation, a
protective mechanism intrinsic to and mediated by the
AR N-terminal domain, and for AR dimerization and
transcriptional activation (Fig. 1) [5, 6]. Recent evidence
suggests that AR forms a head-to-toe (antiparallel) dimer
in its activated state [7]. Androgen binding also displaces
receptor associated proteins, such as the 90-kDa heat
shock proteins [8–10], which may relieve constraints on
receptor dimerization and/or subsequent DNA binding
[10, 11]. Once imported to the nucleus, AR binds andro-
gen response element DNA regulatory sequences within
or flanking androgen responsive genes and activates tran-
scription of those genes.
Antihormone binding induces a conformation that dif-
fers from that imposed by agonist binding (Fig. 1) as
originally suggested by studies on the estrogen receptor
(ER) [12–14]. Limited protease digestion studies demon-
Fig. 1 Schematic illustration
depicting AR androgen binding
and activation and the effects of
antiandrogens. Unliganded AR
is susceptible to rapid degrada-
tion by proteases. High-affinity
androgen binding induces a
conformational change that is
requisite for AR stabilization,
dimerization, and transcription-
al activation. Two androgen-
bound AR monomers (red and
gray) form a head-to-toe, anti-
parallel, dimer, but it has not
been established whether AR
dimerization occurs in solution
or requires binding to androgen
response element DNA (yellow
ovals). Antiandrogens bind
AR with moderate affinity and
induce a conformation that fails
to protect AR against degrada-
tion or is compatible with AR
dimerization and/or ARE-DNA
binding. Antiandrogens inhibit
transcription by preventing AR
binding to androgen response
element DNA&/fig.c:
Sex differentiation:
role of androgens and the androgen receptor
Alterations in sex differentiation and development are
not lethal and therefore provide information regarding
the role of sex steroids in mammalian reproductive de-
velopment [39, 40]. Genetic sex is determined at fertili-
zation by the presence of the Y chromosome that directs
differentiation of the indifferent gonads into testes. Prior
to gonadal sex differentiation the embryo can potentially
develop into the male or female phenotype. Following
gonadal sex differentiation testicular androgen secretion
induces differentiation of the male internal ducts and ex-
ternal genitalia resulting in the male phenotype. In the
human embryo the onset of testosterone synthesis by the
testis occurs approximately 65 days after fertilization
and induces differentiation of the wolffian duct into epi-
didymides, vas deferens, and seminal vesicles. The tes-
tosterone metabolite DHT induces development of the
prostate and male external genitalia. In the absence of
testosterone the female phenotype is expressed indepen-
dently of the presence of an ovary.
Immunostaining and reverse transcriptase polymerase
chain reaction indicates the presence of AR in most tis-
sues with highest concentrations in the male reproductive
tract [41]. A single AR gene located on the long arm of
the X chromosome at Xq11-12 [42–44] encodes a single
AR protein of apparent molecular weight 110–114 kDa,
comprising 910–919 amino acids [45–47]. The variabili-
ty in length is due to the polymorphic glutamine (CAG)
repeat in the N-terminal domain [48–52]. AR mediates
sex differentiation in the developing fetus and androgen
action in postembryonic life, as immunoblots prepared
from urogenital tract tissues of gestational day 17 male
and female rats recognize the same 110-kDa protein
characteristic of AR from adult tissues [53]. In addition,
the ligand-binding characteristics of AR isolated from
the embryonic urogenital tract are the same as those de-
termined from mature adult reproductive tract tissues
[53]. The enhanced sensitivity of the developing male fe-
tus to the antiandrogenic effects of environmental chemi-
cals may result from reduced levels of competing endog-
enous androgens in the fetal male compared to the adult.
Since basic mechanisms of sex differentiation are sim-
ilar in mammals, chemicals that affect reproductive devel-
opment in rodents and other mammals should be consid-
ered potential human reproductive toxicants as well. Ob-
viously, species differences in the health effects of these
chemicals can occur depending on the ability of each spe-
cies to metabolize and/or distribute the test chemical; this
is also true among individuals of different ages within a
species. In this regard, chemical exposure during sex dif-
ferentiation is of particular concern for several reasons.
First, development of the reproductive system often is
sensitive to low-dose chemical effects. Second, while
chemical exposure may be transient, the effects, including
reproductive capacity, are irreversible. Third, functional
alterations are often not discovered until after puberty or
200
strate conformational differences induced by agonists
and antagonists for AR [15, 16], ER [17, 18], and the
progesterone receptor (PR) [19, 20]. Even single amino
acid changes in steroid receptor ligand binding domains
can alter the structure/function relationships of agonist
and antagonist. For example, substitution of alanine for
threonine 877 [21–23] in the LNCaP cell AR alters the
specificity of ligand binding such that low concentrations
of estrogens [24, 25], progestins [26], and antiandrogens
such as hydroxyflutamide, cyproterone acetate, and nil-
utamide [23, 27–29] have agonist activity and promote
LNCaP cell growth and secretion of androgen-dependent
proteins [26, 30]. A single base mutation in ER reduces
agonist binding without altering binding affinity for the
antihormone tamoxifen [31, 32]. A single amino acid
mutation in PR prevents RU486 binding but has little ef-
fect on progesterone binding [33]. While it is clear that
different ligands induce different receptor conforma-
tions, it remains to be established how these structural al-
terations influence receptor dimerization, DNA binding,
and/or transcriptional activation.
Prevalent models for antihormone action include two
mechanisms of transcriptional inhibition that reflect the
ability of receptors to bind DNA. Type I antagonists bind
the receptor but prevent receptor DNA binding, whereas
type II antagonists induce DNA binding but fail to initi-
ate transcription [34]. Insight into these mechanisms has
been provided by studies with PR. PR binds response el-
ement DNA whether associated with agonist R5020 or
antagonist RU486 [35, 36], but apparently with altered
conformation as differences were observed in the migra-
tion of PR DNA complexes on sucrose gradients and in
DNA mobility shift assays [35]. Migration of ligand-re-
ceptor complexes in DNA mobility shift assays differ de-
pending on agonist or antagonist binding, supporting a
role for ligand-induced conformational changes in the re-
ceptor when bound to DNA. For human PR antagonist-
bound complexes migrate more rapidly than agonist
complexes, and there is evidence that binding of two dif-
ferent ligands in the same receptor dimer (mixed-ligand
dimer) alters PR affinity for DNA [36, 37]. Type I antag-
onists, then, may form mixed-ligand dimers with little to
no affinity for DNA.
The environmental antiandrogens identified to date
exhibit low to moderate affinity for AR and act as type I
antagonists by preventing AR DNA binding (Fig. 1) [2,
38]. While we have previously suggested that the forma-
tion of mixed-ligand dimers is required for AR antago-
nism [38], the specific mechanisms responsible for inhi-
bition of AR DNA binding may include, but are not lim-
ited to: (a) increased AR degradation via an inappropri-
ate receptor conformation or inability of rapidly dissoci-
ating ligands to stabilize AR [5, 27], (b) the AR dimer-
ization interface for agonist- and antagonist-bound AR
are incompatible, and no mixed-ligand dimers form, or
(c) mixed-ligand AR dimers fail to bind DNA due to an
inappropriate dimer conformation [7] or an inability to
release receptor-associated proteins requisite for subse-
quent DNA binding [15, 16].
201
later in life, leading to underestimates of chemical in-
duced effects on reproductive development. Lastly, devel-
opmental abnormalities cannot be predicted from chemi-
cal exposures in adult animals as adults are fully differen-
tiated and can therefore tolerate a greater chemical insult.
Developmental reproductive toxicity data are therefore
often required for the assessment of noncancer health ef-
fects of endocrine-disrupting chemicals.
Environmental antiandrogens:
developmental effects and molecular mechanisms
Vinclozolin
The fungicide vinclozolin (3-(3,5-dichlorophenyl)-5-
methyl-5-vinyl-oxazolidine-2,4-dione; Fig. 2) has antian-
drogen activity and alters sex differentiation in male rats
by inhibition of AR-mediated gene activation [38, 54]. In
male rat offspring perinatal exposure to vinclozolin caus-
es hypospadias, ectopic testes, vaginal pouch formation,
agenesis of the ventral prostate, and nipple retention,
whereas female offspring are phenotypically normal
[55]. Exposure to 50 mg vinclozolin/kg per day during
development (gestational day 14 to postnatal day 3)
causes infertility and reduced ejaculated sperm counts in
adult male offspring primarily due to the presence of se-
vere hypospadias. Concentrations as low as 12 mg/kg per
day permanently reduce ventral prostate weight follow-
ing developmental exposure. In contrast, fertility is unaf-
fected in adult male rats after prolonged high-dose expo-
sure (100 mg/kg per day for 25 weeks) [56]. The results
indicate that the developing fetus is particularly sensitive
to endocrine disruptors such as vinclozolin, which can
produce malformations at doses that have little or no re-
productive effect in adult males. The extent to which the
human population is exposed to vinclozolin and/or its ac-
tive metabolites either directly or via the food chain is
not known.
The molecular mechanisms responsible for the antian-
drogenic effects of vinclozolin were recently elucidated
[38, 54]. The two primary metabolites of vinclozolin – 2-
[[(3, 5-dichlorophenyl)-carbamoyl] oxy] -2-methyl-3-
butenoic acid (M1) and 3′,5′-dichloro-2-hydroxy-2-
methylbut-3-enanlide (M2) – compete for AR androgen
binding and inhibit DHT-induced transcriptional activa-
tion (Fig. 3) by blocking AR binding to androgen re-
sponse element DNA [38]. The parent compound vi-
nclozolin, on the other hand, is a poor inhibitor [54]. The
structural similarity of M1 and M2 to the potent antian-
drogen hydroxyflutamide is illustrated in Fig. 2. At high
concentrations in the absence of DHT, M2 targets AR to
the nucleus and acts as an agonist. These diverse effects
of M2 led to the hypothesis that mixed-ligand AR di-
mers, i.e., two different ligands bound in the AR dimer,
are functionally antagonistic whereas same ligand dimers
act as agonists [38]. A similar hypothesis was recently
suggested for the progesterone receptor [57].
DDE
Another potent environmental antiandrogen is DDE, a
DDT metabolite that bioaccumulates in the environment
Fig. 2 Structural formulas of vinclozolin, vinclozolin metabolites
M1 and M2, DHT, DDE, and hydroxyflutamide&/fig.c:
Fig. 3 Inhibitory effects of increasing concentrations of vinclozolin
and metabolites M1 and M2 on AR-mediated gene transcription.
Transcriptional activity was determined in CV1 cells by transiently
transfecting human AR expression vector pCMVhAR and a mouse
mammary tumor virus promoter-luciferase reporter vector. Cells
were treated with 0.1 nM DHT with or without the indicated con-
centrations of competing ligands. Luciferase optical units are shown
with standard error, and fold induction is indicated (bottom) relative
to the activity determined in the absence of 0.1 nM DHT. p5 repre-
sents results when the cells were cotransfected with the luciferase
reporter vector and the parent AR expression vector pCMV5, which
lacks AR sequence. (Modified from [38])&/fig.c:
Permixon
The natural plant product permixon that has antiandro-
genic activity derives from a liposterolic extract from the
fruit of the American dwarf palm tree, Serenoa repens B,
native to Florida [70]. In clinical trials palm tree extracts
improved the symptoms of benign prostatic hyperplasia,
including dysuria, nocturia, and poor urinary flow [71].
The antiandrogenic effects of permixon are mediated by
direct action through the AR (IC50=367 µg/ml) and by
inhibition of 5-reductase activity (IC
50
=88 µg/ml) [70].
Effects of permixon on the development of the male or
female reproductive systems have not been reported.
Clinical implications
The increased incidence of idiopathic hypospadias defor-
mity (incomplete fusion of the urethral folds of the penis
causing the urethral opening to be displaced from the
end of the penis) and cryptorchidism (undescended tes-
tes) in the newborn male population raises the possibility
that environmental antiandrogens are detrimental to nor-
mal male genital development in utero [72, 73]. The
most common birth defects observed in humans resulting
from inhibition of AR action are alterations in the devel-
opment of the male external genitalia [74]. During the
first 12 weeks of gestation, a period critical in sex differ-
entiation of the human external genitalia [75], fetal an-
drogens induce ventral folding and fusion of the urethral
folds to form the penis and fusion of the labioscrotal
folds to form the scrotum. In the second and third trimes-
ters androgen-dependent growth of these structures con-
tinues. Environmental antiandrogens could potentially
cause male pseudohermaphroditism (incomplete mascu-
linization of the male fetus) to differing degrees depend-
ing to the time of exposure and the potency of the chemi-
cal [76]. This is of particular concern given previous re-
ports suggesting an increase in the incidence of andro-
gen-dependent human male reproductive tract abnormal-
ities in the general population [72, 73].
Environmental chemicals with antiandrogen activity
may contribute to the increasing incidence of isolated
hypospadias in the human population, as this anomaly is
not frequently associated with mutations in the AR cod-
ing sequence [77, 78], with steroid 5-reductase (type 2)
deficiency [79] or with decreased levels of AR [80]. Hy-
pospadias has been linked to fetal exposure to estrogenic
chemicals in the first trimester [81]; however, the exter-
nal genitalia of the human male fetus may lack ER [75].
As estrogens bind AR with moderate affinity [82], the
increased incidence of hypospadias following develop-
mental exposure to estrogenic chemicals may be mediat-
ed at the level of the AR.
Another clinical problem that environmental antian-
drogens may exacerbate relates to the increasing inci-
dence of breast cancer among females. Androgens are
naturally antiestrogenic in breast tissue and suppress the
growth of normal as well as breast cancer cells. Environ-
202
[2]. DDE is found in food [58, 59] and in human body
fat [60, 61], has a clearance half-life in the body and the
environment of over 65 years, and comprises 50–80% of
the total DDT-derived residues in human breast milk mo-
bilized to the suckling infant [62–64]. When adminis-
tered to pregnant rats from gestational day 14-18, DDE
(100 mg/kg per day) reduces anogenital distance and
causes retention of thoracic nipples in male progeny [2],
both of which are indicators of prenatal antiandrogen ex-
posure [65]. DDE binds AR in vitro and inhibits DHT-
induced transcriptional activation with a potency similar
to that of the antiandrogenic drug hydroxyflutamide
(Fig. 4) [2].
Although DDT use was banned in the United States in
1973, it persists in the environment and continues to be
used in other countries. Recent evidence of a global dis-
tillation effect for some organochlorine pollutants indi-
cates migration through the atmosphere from warmer to
colder latitudes [66]. Median levels of DDE measured in
serum (12.6 ppb) and placental tissues (6.8 ppb) from
women in the United States within the past 10 years are
below those required to inhibit androgen action in vitro;
however, maximum levels in certain populations exceed
these concentrations (e.g., 180 ppb in serum and 74 ppb
in placenta) [64], presumably due to continous low level
exposure and slow metabolic clearance. In South Africa
Bouwmann et al. [67] report median DDT/DDE serum
levels from individuals living in dwellings treated with
DDT for malaria control of 140.9 ppb and 475 ppb in
breast milk, leading them to postulate a risk to infants
[68]. In the mid-1960s when DDT was in use in the
United States, high concentrations of DDE were mea-
sured in tissues from stillborn infants in Atlanta, Georgia
(650 ppb in brain, 850 ppb in lungs, 2740 ppb in heart,
980 ppb in liver, 3570 ppb in kidney, and 860 ppb in
spleen) [69]. Although the in vivo cellular concentration
of DDE is not known, reports suggest that human DDE
levels can exceed those that inhibit human AR transcrip-
tional activation in vitro.
Fig. 4 Transcriptional inhibitory effects of increasing concentra-
tions of DDE and the potent antiandrogen hydroxyflutamide. Tran-
scriptional activity was determined as described in Fig. 3. (Modi-
fied from [2, 38])
mental antiandrogens could interfere with this normal
androgen-induced growth suppression leading to hyper-
trophy. The suppresive effects of androgen on breast cell
proliferation are illustrated by feminine breast develop-
ment in genetic males with the androgen insensitivity
syndrome resulting from a mutation in the AR gene [83,
84]. In the absence of a response to circulating androgen,
breasts develop in these subjects due to the presence of
endogenous estrogens in adult males. Although the
mechanisms by which androgens inhibit breast growth
and development are poorly understood, AR is expressed
in 85% of breast cancers [85] and androgen therapy ben-
efits 20–50% of advanced breast cancer patients [86].
Furthermore, early onset of breast cancer is correlated
with low urinary and serum androgen levels [87]. Physi-
ological levels of DHT reduce estradiol formation in ZR-
75-1 human breast cancer cells [88] and suppress ER
mRNA production [89]. Androgens inhibit ER-induced
PR synthesis in MCF-7 human breast cancer cells [90]
and the proliferation of MFM-223 mammary carcinoma
cells [91]. Many of these effects are inhibited by hy-
droxyflutamide, suggesting a direct effect mediated
through the AR. It also is of interest that a higher risk of
breast cancer is reported in women with elevated serum
and fat levels of DDE or polybrominated biphenyl, a
compound structurally related to DDT [92–94], support-
ing the hypothesis that environmental antiandrogens con-
tribute to the increasing incidence of breast cancer devel-
opment in women. A fourfold increase in relative risk of
breast cancer was observed with elevation of serum DDE
from 2 to 19 ng/ml [93]; however, these correlations
have not been universally accepted [95, 96].
Finally, several groups have reported a 2.1% per year
decline in the sperm count of specific adult human male
populations from 89 million/ml (1973) to 60 million/ml
(1992; P<0.001) and an increase in sperm malformations
[97–102]. In these areas the year of birth is a major pre-
dictor of sperm concentration, suggesting that the de-
cline could result from neonatal and/or pubertal exposure
to toxicants perhaps with antiandrogenic activity. As an-
drogens are necessary to initiate spermatogenesis at pu-
berty and maintain spermatogenesis in the adult [103], it
is conceivable that environmental antiandrogens contrib-
uted to the decline, although chronic high-level exposure
likely would be required. The decline in human sperm
counts and/or sperm quality is another highly conten-
tious topic [104, 105].
Testing strategies
Multigenerational reproductive toxicology test protocols
encompass both developmental chemical exposure mea-
surements and continuous animal health monitoring
[106]. These protocols do not currently require endocrine
data or even bioassay measurements for hormone activi-
ty [107]. Laboratories at the United States Environmen-
tal Protection Agency have developed multigenerational
test guidelines that include measures of endocrine func-
tion such as pubertal landmarks, estrous cyclicity, and re-
productive organ weights in an Alternate Reproduction
Test protocol [108, 109]. Because these tests are labor
intensive, require numerous animals, take more than a
year to complete, and are expensive, shorter term test
protocols are sought to identify endocrine disrupting
chemicals and to elucidate their mechanisms of endo-
crine toxicity. It is doubtful that single in vivo or in vitro
tests will adequately assess endocrine disrupting activity.
For example, screening chemicals in vitro for estrogenic-
ity may fail to identify compounds that act through other
steroid hormone receptors, alter steroid hormone biosyn-
thesis, transport, or degradation, or require metabolic ac-
tivation. Such in vitro estrogenicity tests alone would fail
to predict the antiandrogenic in vivo effects induced by
DDE or vinclozolin.
We propose combined in vivo and in vitro test strate-
gies to screen chemicals for reproductive toxicity. In vivo
tests identify chemicals with endocrine disrupting activi-
ty, whereas in vitro tests reveal the effective chemical or
metabolite and provide information regarding the bio-
chemical mechanism. Given this information, human
susceptibility and risk assessment issues can be subse-
quently addressed. Dosing newborn animals with toxic-
ants has been a successful strategy in detecting estroge-
nicity and antiandrogenicity. The age and weight at pu-
berty, weights of the reproductive organs, and serum hor-
mone levels are measured from days 22 to 50 in male
and female rats that were treated since birth. This ap-
proach has revealed the toxic effects of chlordecone, me-
thoxychlor, vinclozolin, and DDE.
This approach of Kelce et al. [2, 54] and Gray et al.
[55] combines in vivo studies to identify endocrine dis-
rupting chemicals with in vitro studies to characterize
the molecular mechanism of action. For example, studies
with vinclozolin by the manufacturer, BASF-Agrochemi-
cals, suggested that vinclozolin has antiandrogenic activ-
ity. Using the above in vivo protocol, vinclozolin admin-
istered for 30 days starting at weaning delayed puberty
reduced the weights of the sex accessory glands and in-
creased serum testosterone and luteinizing hormone lev-
els, consistent with the endocrine profile of an antiandro-
gen. In a developmental study vinclozolin caused altera-
tions in male rat sex differentiation such as reduced ano-
genital distance, cleft phallus, hypospadias, ectopic tes-
tes, and retained thoracic nipples in male offspring [55],
all indicative of antiandrogen activity. In vivo experi-
ments such as these helped to identify vinclozolin as an
antiandrogen endocrine disruptor.
Subsequent in vitro studies demonstrated that al-
though the parent chemical, vinclozolin, bound AR
weakly, its two primary hydrolysis products were strong-
er androgen antagonists [54]. Maternal serum concentra-
tions of these metabolites were at levels near the K
i
for
inhibition of androgen binding, suggesting that the de-
velopmental toxicity of vinclozolin was mediated by its
hydrolysis products, M1 and M2 [54]. Molecular studies
determined that the mechanism of inhibition of andro-
gen-induced transcriptional activation was inhibition of
203
AR DNA binding [38]. Within a relatively short period
and with a limited number of animals, the antiandrogenic
activity of vinclozolin was identified, its adverse devel-
opmental effects characterized, and the biochemical and
molecular mechanism established. Thus, combining in
vivo and in vitro test strategies is effective in the identifi-
cation and characterization of environmental endocrine
disruptors.
The above testing strategy was used to identify over
25 environmental chemicals predicted to have an affinity
for AR. Those structures shown biochemically to interact
with AR were introduced into a computer model together
with known androgen agonists and antagonists to devel-
op a three-dimensional quantitative structure-activity re-
lationship paradigm that predicts AR binding affinity
solely from chemical structure, taking into account steric
and electrostatic properties [110]. The model is being
used to search structural databases for potential androgen
agonists and antagonists. Chemicals identified using the
computer are subsequently examined for AR binding ac-
tivity and induction or inhibition, respectively, of andro-
gen-dependent transcriptional activity. Chemicals that al-
ter androgen action in vitro are further tested in vivo fol-
lowing the above strategy. To date the computer model
has identified several hundred chemicals with the poten-
tial to bind AR. Empirical testing has begun and already
several novel androgen antagonists have been identified
[110]. The impact of these environmental chemicals on
the development and health of wildlife and humans re-
mains to be appreciated.
&p.2:Acknowledgements This manuscript has been reviewed in accor-
dance with the policy of the National Health and Environmental
Effects Research Laboratory, United States Environmental Protec-
tion Agency, and approved for publication. Approval does not sig-
nify that the contents necessarily reflect the views and policies of
the Agency, nor does mention of trade names or commercial prod-
ucts constitute endorsement or recommendation for use. The au-
thors thank Drs. L Earl Gray, Jr, Gary Klinefelter, and Tom Wiese
for their critical review of the manuscript. The authors also grate-
fully acknowledge Dr. Gray for many helpful discussions regard-
ing the development and implementation of in vivo/in vitro testing
strategies for environmental endocrine disrupting chemicals.
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