TOXICOLOGICAL SCIENCES 120(S1), S93–S108 (2011)
doi:10.1093/toxsci/kfq329
Advance Access publication October 25, 2010
Endocrine Disruption: Historical Perspectives and Its Impact on the
Future of Toxicology Testing
Mary Sue Marty,1 Edward W. Carney, and Justin Craig Rowlands
Toxicology and Environmental Research and Consulting, The Dow Chemical Company, Midland, Michigan
1
To whom correspondence should be addressed at Toxicology and Environmental Research and Consulting, The Dow Chemical Company, 1803 Building,
Midland, MI 48674. Fax: (989) 638-9863. E-mail: [email protected].
Received August 13, 2010; accepted October 15, 2010
The endocrine system is one of the body’s major
homeostatic control systems whose aim is to maintain normal
functions and development in the face of a constantly changing
environment. Working in tandem with the nervous system,
which mainly is responsible for rapid and immediate responses,
the approximately 30 different glands comprising the endocrine
system tend to act in a slower and more sustained manner to
regulate processes as diverse as the female reproductive cycle,
bone growth, cell proliferation, and psychosocial behavior.
Multiple endocrine glands also work in concert with one
another to form complex feedback loops, which tightly regulate
critical physiological processes.
Like all homeostatic control systems, the capacity to
maintain physiological parameters within normal bounds is
finite, and when this capacity is exceeded by chemical
exposures, drugs, or environmental stressors, adverse
consequences can ensue. In the mid-1990’s, concerns about
the potential for environmental chemicals, drugs, and other
stressors to alter endocrine physiology rapidly mounted and
received a great deal of attention within the toxicology
community as well as in the public media. These circumstances
spurred a flurry of research on mechanisms of toxicity, new
questions about how to deal with cumulative exposures to
endocrine-active compounds (EACs) from man-made, dietary,
and endogenous sources, and the creation of major chemical
screening and testing programs focused on endocrine-mediated
modes of toxic action.
In this review, we will trace the path of knowledge regarding
the effects of exogenous agents on the endocrine system, the
events which gave rise to the explosion of interest in this topic
in the mid-1990’s, and how it led to a fundamentally different
paradigm in toxicity testing as well as a great deal of basic
research on mechanisms of endocrine-mediated toxicity. We
first cover the history of testing for endocrine-mediated
toxicity, followed by coverage of key historical events in
basic science, and then bring the two together to discuss
implications of the science on testing both now and for the
future.
EARLY HISTORY OF TOXICITY TESTING FOR
ENDOCRINE-MEDIATED EFFECTS
A casual observer might easily be led to believe that the
ability of exogenous substances to interfere with the endocrine
system was unknown until relatively recently, but in fact, this is
far from true. Even the ancients were familiar with the actions
of various herbal preparations to modulate what we now know
to be hormonally regulated processes. For centuries, farmers
have observed reproductive problems in female sheep and
cows grazing on pastures rich in certain clover species, which
later were found to contain estrogenic compounds such as
coumestrol (Adams, 1995). As early as the 1930’s, the ability
of both natural and synthetic chemicals to interact with
endogenous hormone receptors was already well established
(Schueler, 1946; Sluczewski and Roth, 1948; Walker and
Janney, 1930).
Starting around the 1940’s, steroidal compounds began to be
used in the livestock industry to modulate reproductive cycles
and to enhance rate and efficiency of body weight gain, while
compounds such as ethinyl estradiol, mestranol, norethynodrel,
and megestrol acetate were being developed as contraceptive
agents and to treat specific disease conditions in humans
(Newburgh, 1975). In parallel with these developments,
toxicologists began to investigate these compounds for potential adverse effects.
Diethylstilbestrol (DES), a ‘‘synthetic estrogen’’ with similar
potency to 17b-estradiol, was first prescribed in 1938 to
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MARTY, CARNEY, AND ROWLANDS
prevent miscarriage and premature births. This drug was
particularly significant as the daughters of women who took
DES were found to have an increased incidence of clear-cell
carcinoma, a rare vaginal cancer. As a result, in 1971, the Food
and Drug Administration (FDA) advised physicians to stop
prescribing DES to pregnant women. Further research has
revealed an increased risk for breast cancer in DES mothers
and also possible links with reproductive tract abnormalities in
DES sons and daughters (Giusti et al., 1995). It has been
estimated that 5–10 million Americans received DES during
pregnancy or were exposed to the drug in utero.
Other episodes of endocrine-mediated toxicity also occurred
via unintentional exposure to hormonally active compounds,
such as polychlorinated biphenyls (PCBs) and polychlorinated
dibenzofurans (PCDFs). Some members of this large family of
structurally related congeners can act on the estrogen and
arylhydrocarbon (Ah) receptor systems and/or modulate
thyroid hormone function. In Japan (1968) and Taiwan
(1979), a few thousand women were exposed to PCBs and
their pyrolysis products (PCDFs) following consumption of
contaminated rice oil. The offspring of these mothers exposed
to PCBs and PCDFs tended to be smaller at birth and
exhibited delays in neurological development (Aoki, 2001).
Subsequent studies on PCB exposure at lower levels have
alleged other adverse effects, although these data remain
somewhat controversial.
In the environment, the presence of natural hormones in
wastewater treatment outfalls in the United States was noted as
early as 1965 (Stumm-Zollinger and Fair, 1965), and this work
expanded in 1970 to include synthetic estrogens used as birth
control agents (Tabak and Bunch, 1970). By the mid-1970’s,
scientists began to detect other pharmaceutical agents near
wastewater outlets in the United States (Garrison and Pope,
1975; Hignite and Azarnoff, 1977). Concerns about dichlorodiphenyltrichloroethane (DDT) in the environment were
highly publicized in Rachel Carson’s book, Silent Spring,
although a specific link to endocrine-mediated toxicity was yet
to come. Despite the aforementioned events in both wildlife
and humans, endocrine-mediated toxicity did not garner any
special attention above and beyond other mechanisms of
toxicity known at the time. Instead, it tended to be regarded as
just one of many potential mechanisms that could lead to
certain adverse outcomes, such as reproductive toxicity,
developmental toxicity, or cancer.
THE BIRTH OF ‘‘ENDOCRINE DISRUPTORS’’
This mindset changed rather abruptly in the mid-1990’s
following a series of publications suggesting that pesticides and
other man-made chemicals were threatening the reproductive
capability and intelligence of future generations of humans and
wildlife (Colborn et al., 1996, 1993; Hunter and Kelsey, 1993).
Of particular impact was a book entitled Our Stolen Future, Are
We Threatening Our Fertility, Intelligence and Survival?—
A Scientific Detective Story written by Colborn et al. (1996)
with a forward penned by then Vice President Al Gore. These
and other authors proposed that many EACs elicited effects at
doses far lower than toxicities caused by other modes of
action (MOAs) and thus required special regulation. Also
around the same time were a number of more specific reports
alleging associations between endocrine-disrupting chemicals
and declining sperm counts (Carlsen et al., 1992; Sharpe and
Skakkebaek, 1993), increases in breast cancer (Davis et al.,
1993), cryptorchism and hypospadias (Barlow et al., 1999),
and developmental abnormalities in wildlife species (Sumpter,
1995). The level of concern spiked even higher following a
1996 report claiming > 1000-fold level of synergy among
a mixture of four weakly estrogenic agents as assessed in a
yeast estrogen receptor (ER) transcriptional activation assay
(Arnold et al., 1996). Ironically, the latter results could not be
reproduced by others (Ashby et al., 1997; Ramamoorthy
et al., 1997) and a National Institutes of Health inquiry
revealed intentional falsification of the results leading to
retraction of the original report, which had been published in
Science (McLachlan, 1997).
Despite uncertainties about the scientific validity of these
reports and the true extent of the problem, both the scientific
community and public reacted as if this were an approaching
tsunami. Media coverage of what would previously have been
considered an esoteric science topic became mainstream
headline news. Prominent government and scientific organizations such as the World Health Organization, International
Union of Pure and Applied Chemistry and the U.S. National
Research Council reacted by gathering endocrinologists,
toxicologists, and other public health professionals from
around the world to capture their knowledge of the state-ofthe-science and to recommend research to address the issue.
These efforts are documented in a series of comprehensive
review papers (IPCS, 2002; Lintelmann et al., 2003; NRC,
1999), which also helped spur a plethora of research on
endocrine-mediated toxicity. The IPCS (2002) report was
particularly significant in that it described a structured weightof-evidence approach to guide the application of new research
for addressing causal relationships between exposure to
EACs and adverse health outcomes. In fact, application of
weight-of-evidence approaches has failed to provide convincing support for many of the early claims that initially
sparked the issue, such as alleged links between EACs and
declining sperm counts (Fisch, 2008) and breast cancer
(summarized at: http://epi.grants.cancer.gov/LIBCSP/ ). Reported effects in aquatic species have pointed more toward
natural and synthetic hormones (e.g., birth control pills) as
causative agents, more so than industrial chemicals as
originally suggested (Jobling et al. 2006). Nonetheless, the
term endocrine disruptor has been adopted in the vernacular
of toxicologists, public health professionals, and even the
general public and appears here to stay.
ENDOCRINE-MEDIATED TOXICITY
Somewhat unwittingly, the new emphasis on endocrine
disruption also marked the beginning of a fundamental change
away from the traditional ‘‘outside-in’’ approach to toxicology,
in which adverse outcomes are identified first followed by
mechanistic research to understand these effects. With
endocrine disruption came a reversed approach in which
identification of mechanism came first followed by research to
define possible consequences of activating these mechanisms.
The implications of this change are quite significant and are
discussed later in this review.
GENESIS OF ENDOCRINE SCREENING AND TESTING
At around the same time that these alarming reports were
emerging, the U.S. Safe Drinking Water Act was being updated
while the Food Quality Protection Act was being created to
enhance protection of children and other vulnerable subpopulations. This temporal convergence led to a 1996 congressional
mandate requiring the Environmental Protection Agency (EPA) to
implement a screening and testing program to detect endocrine
disruptors. Despite the nascent state of the science at the time,
Congress gave EPA just 2 years to establish the new program. To
assist EPA, an advisory committee known as the Endocrine
Disruptor Screening and Testing Advisory Committee (EDSTAC)
was formed to evaluate protocols for approximately 50 different
endocrine assays and select a subset to comprise a ‘‘Tier I’’
screening battery for biological activity involving the estrogen,
androgen, and thyroid (EAT) hormone systems.
Although limiting screening to these three hormone systems
might seem arbitrary, these three appeared to have the broadest
range of potential effects as well as the greatest number of
ligands with which they interact (Blair et al., 2000).
Compounds that can bind steroid hormone receptors include
industrial chemicals such as bisphenol A (BPA), certain
alkylphenol ethoxylates, phthalates, parabens, benzophenones,
and environmental contaminants like DDT, methoxychlor,
some PCBs, dioxins and furans. There also are many
pharmaceuticals purposely designed to have hormonal activity,
such as DES, contraceptive agents, and the selective ERmodulators (SERMs) used in the treatment of diseases such as
osteoporosis, as well as numerous plant-derived compounds
such as genistein, daidzein, naringenin, genistin, daidzin,
glycitein, and puerarin (Boue et al., 2003) and resveratrol, a
constituent of red wine (Klinge et al., 2003). Finally, exposure
to synthetic and plant-derived estrogens occurs against a
background of endogenous estrogens, such as estradiol-17b
(E2), estrone (E1), and estriol (E3), the levels of which in
blood and tissues can vary dramatically depending on gender,
age, and physiological state (e.g., pregnancy, pre-, or postmenopause).
Beyond need to cover so many different chemistries and
sources, the challenge laid before EDSTAC was further
complicated by the desire to screen for both receptor agonists
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and antagonists, as well as compounds which act indirectly.
Indirect mechanisms are plentiful and include the inhibition of
steroid biosynthesis or modulation of the hypothalamicpituitary-gonadal (HPG) axis. An even more complex example
begins with activation of the Ah receptor (AhR), which can
ultimately oppose estrogen signaling.
Ironically, some compounds bearing the very properties
that are the targets of screening can actually be beneficial. Studies
of endocrine-disrupting AhR ligands have led to the development
of selective AhR modulators, such as ring-substituted
diindolylmethanes (DIMs), for chemotherapy of breast,
endometrial, pancreatic, and prostate cancer (Safe et al.,
2008). Substituted furans and indoles have been licensed by a
pharmaceutical company as antitumor development compounds
(Safe, personal communication). Some of the DIMs have been
further modified into anticarcinogenic AhR-inactive analogs,
which activate orphan nuclear receptors that induce genes
associated with growth inhibition and cancer cell death.
EPA’S ENDOCRINE DISRUPTOR SCREENING PROGRAM
TIER I ASSAYS
In 1998, EDSTAC released its report outlining a screening
and testing program to assess the potential of compounds to
interact with the endocrine system. EDSTAC recommended
that a tiered approach be adopted to examine endocrine activity
with a focus on the potential for compounds to interact with the
EAT hormone systems. The Tier I assays were designed to
screen for potential interaction with these endocrine systems
and designed to be an efficient screening process to screen
a large number of chemicals for endocrine activity. In Tier I,
EDSTAC opted to prioritize assays with greater sensitivity (but
possibly lower specificity) to avoid false-negative results.
EDSTAC described the importance of using a ‘‘weight-ofevidence’’ approach to evaluate Tier I screening results. Tier II
tests, such as the multigeneration reproductive toxicity study,
were to be used to characterize dose-response relationships,
whether endocrine-mediated adverse effects for EAT occur, as
well as providing data for risk assessment. During this time
period, EPA also took on an extensive revision of the
multigeneration and some other Tier II tests in order to
enhance their sensitivity to endocrine disruptors.
The Tier I assays included in vitro, in vivo mammalian, and
in vivo nonmammalian screening assays. EDSTAC noted that
EACs may alter endocrine function by affecting the availability
of a hormone at a target tissue (e.g., synthesis, transport,
metabolism, HPG/hypothalamic-pituitary-thyroid [HPT] axis
function, etc.) or by affecting the cellular response to a hormone
(e.g., receptor binding). The original Tier I battery proposed by
EDSTAC (EDSTAC, 1998) included the following: (1) ER
binding/transactivation assay, (2) androgen receptor (AR)
binding/transactivation assay, (3) steroidogenesis assay with
minced testis, (4) rodent 3-day uterotrophic assay (sc route),
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TABLE 1
U.S. EPA EDSP Tier I Assays
In vitro assays
ER binding—rat uterine cytosol
ER—(hERa) transcriptional activation—human cell line (HeLa-9903)
AR binding—rat prostate cytosol
Steroidogenesis—human cell line (H295R)
Aromatase—human recombinant microsomes
In vivo assays
Uterotrophic (rat)
Hershberger (rat)
Pubertal female (rat)
Pubertal male (rat)
Amphibian metamorphosis (frog)
Fish short-term reproduction
Note.http://www.epa.gov/scipoly/oscpendo/pubs/assayvalidation/
tier1battery.htm#assays.
(5) rodent 20-day pubertal female assay with thyroid, (6) rodent
5- to 7-day Hershberger assay, (7) frog metamorphosis assay,
and (8) fish gonadal recrudescence assay. In addition, four
alternate assays were identified: (1) placental aromatase assay,
(2) modified rodent 3-day uterotrophic (ip), (3) rodent 15-day
intact adult male assay with thyroid, and (4) rodent 20-day
thyroid/pubertal male assay. These alternative assays, if
validated, could be used to replace other Tier I assays, namely,
either the intact adult male assay or the pubertal male assay
would replace the pubertal female and Hershberger assays.
Per the Food Quality Protection Act (1996), the endocrine
screening program was required to use ‘‘appropriate validated
test systems’’; consequently, the EPA launched the endocrine
disruptor screening program (EDSP) validation program.
Based on results of methods development and validation work,
some of the proposed Tier I in vitro assays were adopted,
whereas others have undergone some modifications. The final
list of EPA Tier I EDSP screening assays appears in Table 1.
Both the ER binding and ER transactivation assays have been
adopted into Tier I. The ER binding assay measures the ability
of a radiolabeled ligand ([3H]-17b-estradiol) to bind to rat
uterine cytosolic ER in the presence of increasing test material
concentrations. A compound is positive for ER binding if it
effectively competes with [3H]-estradiol binding in a concentrationrelated manner, and an inhibitory concentration that decreases
radioligand binding by 50% (IC50) is obtained. The ER
competitive binding assay cannot distinguish agonists from
antagonists. Test guidelines also were issued for the Stably
Transfected ER Transcriptional Activation assay using HeLa9903 cells (Organization for Economic Co-operation and
Development [OECD] TG 455; OPPTS 890.1300), although
detection of agonists is all that is required. Although having
limited metabolic capability, the ER transactivation assay
reportedly has good concordance with in vitro ER binding and
in vivo uterotrophic responses for active parent compounds
(Takeyoshi, 2006); however, the test guideline criteria for
a positive response (i.e., if a compound achieves a PC10, which
is 10% of the positive control response at 1nM 17b-estradiol)
may result in numerous false-positive results and jeopardize the
predictiveness of this assay.
AR binding using rat prostate cytosolic preparations also is
included in Tier I (OPPTS 890.1150), although an AR
transactivation assay for use in Tier I screening is still being
developed. The AR binding assay shares many of the same
advantages and limitations as the ER binding assay.
To detect compounds affecting steroidogenesis, the minced
testis assay was originally proposed. This assay was designed
to detect enzyme inhibition in the steroid biosynthesis pathway
by measuring decreased testosterone production; however, the
assay was removed from consideration because of high falsepositive and false-negative responses and difficulty assessing
Leydig cell cytotoxicity in the mixed population of cells
present in minced testis preparations (Powlin et al., 1998). The
steroidogenesis assay was replaced with a cell-based system
that uses a human cell line, the H295R cells (Hecker et al.,
2006, 2008). This assay is designed to measure production of
testosterone and 17b-estradiol by the H295R cells in the
presence of varying concentrations of test compound to
identify alterations in steroid biosynthesis. Because of the
limited metabolic capacity of these cells, the assay primarily
examines the ability of the parent material to interfere with
steroidogenesis. A positive result occurs if there is a significant
difference in steroid production in the presence of less than
20% cytotoxicity.
Aside from steroidogenesis alterations, a separate in vitro
assay was developed specifically to identify compounds
affecting aromatase (CYP19), the enzyme that converts
androgens to estrogens. The aromatase assay examines effects
of the parent compound on recombinant human aromatase
activity by measuring the production of heavy water (3H2O)
during the conversion of 3H-androstenedione to estrone.
Compounds that competitively inhibit aromatase activity will
decrease estrone production and consequently 3H2O levels in
a dose-related manner.
Interlaboratory validation has been conducted with these
Tier I in vitro assays to verify each assay’s ability to detect
known EACs operating via relevant MOAs, although the size
of the validation programs (i.e., number of test compounds,
number of laboratories, etc.) has differed by assay. One issue
with the in vitro EDSP assays is that the assays have limited
ability to metabolize substances; thus, if a metabolite is the
EAC, it may not be detected by these assays. Options to
improve metabolic capacity are under consideration. Although
in vitro assays provide useful information on potential MOA
and can help clarify in vivo assay results, previous studies have
suggested that in vivo data should carry more weight than
in vitro data when determining the need for Tier II testing (e.g.,
Charles et al., 2005).
The uterotrophic assay went through an extensive validation
program, which was led by the OECD Endocrine Disruptor
ENDOCRINE-MEDIATED TOXICITY
Testing and Assessment (EDTA) Task Force and the
Validation Management Group (VMG)—Mammalian. This
validation process, led by Japanese representatives on the
EDTA, included a detailed background document prepared by
the OECD (2003) and multiphase intra- and interlaboratory
validation studies involving approximately 20 laboratories to
show transferability, reproducibility, and reliability. This
validation program tested a potent estrogen (ethinyl estradiol),
weak estrogen agonists in both a coded (blinded) and doseresponse manner to evaluate sensitivity, and negative
compounds to verify specificity. This work culminated in
a series of publications (Ashby et al., 2003; Kanno et al.,
2001, 2003a,b; Owens and Köeter, 2003; Owens et al., 2003)
and an OECD test guideline in 2007. The U.S. EPA test
guideline was released in 2009.
The assay uses female rats (or mice) with nonfunctional
HPG axis to yield sensitive and consistent increases in uterine
weight in response to stimulation by estrogenic materials.
Ultimately, the test guideline outlines the application of the
uterotrophic assay for the detection of compounds with
estrogenic activity; use of this assay to identify antiestrogens
is supported by the test guideline, but not required. Overall, the
uterotrophic assay appears to perform well for the identification
of estrogenic compounds, although the ovariectomized model
has been recognized as more specific for identification of
estrogenic activity than the immature model. The immature
model, which has an intact HPG axis, can detect aromatizable
androgens and is more sensitive to body weight–mediated
changes in uterine weights. The alternate uterotrophic assay,
which used ip dosing, was not included in the validation
because of potential localized effects of the test compounds on
uterine tissue.
The Hershberger assay also underwent an extensive multilaboratory validation process, led by U.S. EPA through the
OECD EDTA VMG—mammalian group (Owens et al., 2006,
2007). The Hershberger assay uses castrated adult male rats
with a nonfunctional HPG axis. Although originally proposed
with a 5- to 7-day exposure period, rats are treated with the
test compound for 10 days by oral gavage or sc dosing in
the adopted version of the assay. Increases in accessory sex
tissue (AST) weights (i.e., glans penis, ventral prostate,
seminal vesicles with fluid, Cowper’s glands, and levator
ani-bulbocavernosus muscles) indicate potential androgenicity.
To evaluate antiandrogenic potential, rats are given the
test compound (oral gavage or sc) concurrent with 0.2 or
0.4 mg/kg/day testosterone propionate sc. If the test compound
inhibits the testosterone-induced increase in AST weights, then
the compound has potential antiandrogenic activity or is
a potential 5a-reductase inhibitor. Two or more organ weights
must be significantly altered to yield a positive result. Overall,
the Hershberger assay appears to perform well for the
identification of potential androgens, antiandrogens, and
5a-reductase inhibitors. During the Hershberger validation
studies, a weanling (non-castrated) animal model was included;
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however, the weanling model did not detect weak antiandrogens
as consistently as the castrated adult model (Freyberger and
Schladt, 2009) and therefore was not included in either the EPA
or the OECD test guideline.
Initially, the 20-day female pubertal assay was recommended, and the 20-day male pubertal assay was an alternate EDSP
screen. Prior to initiation of the interlaboratory validation work,
the exposure period for the pubertal male assay was increased
from 20 to 30 days, and ultimately, both assays were adopted
into the Tier I EDSP battery. One of the strengths of the male
and female pubertal assays is that these assays look for changes
in young animals during a dynamic period of endocrine
activity. The pubertal assays are the most apical of the Tier I
in vivo mammalian assays, but these assays are multimodal,
capable of detecting EACs that operate through a variety of
MOAs (i.e., strong and weak (anti)estrogenic and (anti)androgenic compounds, steroid biosynthesis inhibitors, aromatase
inhibitors, 5a-reductase inhibitors, thyroid-active agents, and
compounds affecting the HPG or HPT axis). Some of these
MOAs are not detected by other Tier I in vitro or mammalian
assays; thus, it may be difficult to dismiss a positive result in
the pubertal assays even if there is some suspicion that the
primary MOA is not endocrine mediated.
One limitation of the pubertal assays is that it may be
difficult to determine the MOA for potential EACs. Some
endpoints assessed in these assays (puberty onset, estrous
cycle, and organ weights) can be altered by chemicals that
operate via an endocrine MOA or by nonspecific/systemic
toxicity. During the validation program, some compounds were
not identified for their primary MOA. For example, phenobarbital, which is a weak thyroid-active compound, did not alter
thyroid weights, thyroid histopathology, or serum T4 or thyroid
stimulating hormone levels in the male pubertal assay during
the validation program but rather showed signs of antiandrogenicity, which included delayed preputial separation and
decreased reproductive and AST weights (U.S. EPA, 2007).
There remains some question regarding the specificity of the
pubertal assays. The Interagency Coordinating Committee on
the Validation of Alternative Methods defines specificity as the
proportion of inactive substances that are correctly identified.
During assay validation, 2-chloronitrobenzene (2-CNB) was
selected as a negative control chemical for the male and female
pubertal assays; however, 2-CNB caused thyroid effects,
delayed vaginal opening, and altered adrenal and uterine
weights in females, and delayed preputial separation, decreased
serum testosterone levels, decreased androgen-dependent organ
weights, and had testis histopathological effects in males. The
EPA selected hydroxyatrazine as an alternate negative control
compound and recently reported that it was negative in both the
male and the female pubertal assays (Stoker and Zorrilla,
2010). Only a subset of these data has been published. Once
available, the full data set is likely to provide additional
information on setting maximum tolerated dose levels and
interpreting pubertal assay results (i.e., provide perspective for
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hydroxyatrazine results reported in Laws et al., 2003). With
few negative compounds included in the pubertal assay
validation, it is unclear whether the male and female pubertal
assays are prone to nonspecific effects, which will produce
a high frequency of false-positive results.
One alternative assay proposed as part of the EDSP was the
15-day intact adult male rat assay. Approximately 30 EACs,
across a variety of MOAs, have been evaluated using the
15-day intact male assay including estrogen agonists/antagonists,
androgen agonists/antagonists, steroid biosynthesis inhibitors,
thyroid-active compounds, and prolactin modulators. The
validation program also examined assay specificity by evaluating
the impact of body weight effects on assay endpoints (O’Connor
et al., 2000) and including a negative control compound, allyl
alcohol (O’Connor et al., 2008). The intact male assay was
specifically designed as an MOA screening study, whereby
researchers could compare the ‘‘fingerprint’’ of the organ weight,
histopathology, and hormonal endpoints for an unknown
compound with data generated for a series of positive controls
to help characterize the MOA (O’Connor et al., 2002). There was
some concern regarding the use of numerous serum hormone
measurements in this assay as serum hormone levels are
inherently variable. However, data developed during the
validation process have shown that serum hormone measurements are reliable with the group sizes employed (n ¼ 15 per
dose). Of greater concern for the U.S. EPA was the inability of
the intact male assay to reliably detect weak antiandrogens
(O’Connor et al., 1999), which would require that the assay be
paired with an alternative assay (e.g., Hershberger assay) for Tier
I screening. Because of this weakness, the 15-day intact male
assay was not included in the EPA EDSP.
In addition to the in vitro and in vivo mammalian assays, two
nonmammalian assays also are included in the Tier I EDSP:
the amphibian metamorphosis assay and the 21-day fish
reproduction screen. The amphibian metamorphosis assay
(OPPTS 890.1100) provides an environmental screen to detect
compounds that have the potential to interfere with the HPT
axis. Xenopus laevis tadpoles are exposed for 21 days starting
at Nieuwkoop and Faber stage 51 (Nieuwkoop and Faber,
1994) and examined for mortality, developmental stage, hind
limb length, snout-vent length, wet weight, and thyroid
histopathology. Because the rate of metamorphic development
is controlled by thyroid hormone in X. laevis, developmental
stage (which is determined by the inspection of various
morphological changes associated with metamorphosis) and
hind limb length (which is normalized by snout-vent length) are
used to determine if development is accelerated, asynchronous,
delayed, or unaffected by the test compound after either 7 or
21 days of exposure. Many of the endpoints assessed in this
assay are apical and can be altered by MOAs that are not
specific for thyroid effects. Some endpoints (e.g., snout-vent
length and wet weight) provide information on whether the
test compound alters growth and/or causes systemic toxicity.
If any changes in developmental stage are observed
(i.e., advanced, delayed, or asynchronous development), thyroid
histopathology is an important endpoint to confirm a specific
thyroid effect. To facilitate interpretation, laboratories are
encouraged to include positive control compounds to verify
assay performance and build historical data. Thus, the
amphibian metamorphosis data should be considered in
a weight-of-evidence approach to determine potential thyroid
activity.
Originally, the fish gonadal recrudescence assay was
proposed as an EDSP Tier I screening assay. In this assay,
fish were examined for compound effects on maturation from
a regressed state (recrudescence) in response to an increased
photoperiod and temperature regime. Endpoints included
secondary sex characteristics, ovary and testis weight increases,
gonadosomatic index, gamete maturation, and induction of
vitellogenin. However, feasibility studies indicated that there
was high variability in the gonadal recrudescence response in
continuous spawning species like the fathead minnow (Touart,
2005); thus, the recrudescence assay was not included in the
EDSP. Instead, the U.S. EPA has opted for the 21-day fish
reproduction screen (OPPTS 890.1350) in the Tier I battery as
opposed to the OECD 21-day fish assay (OECD TG 230). This
screen uses a minimum of 96 reproductively mature adult
fathead minnows (alternate models allowed per OECD)
exposed continuously for 21 days to a minimum of three test
concentrations and a control. The EPA 21-day fish reproduction screen includes additional parameters not measured in the
OECD 21-day fish screen, including fecundity, fertility,
gonadal histopathology, and sex steroid concentrations. In
addition, adult survival, secondary sex characteristics (such as
coloration patterns and tubercle scores), reproductive behavior,
vitellogenin, and gonadal-somatic index (GSI) are evaluated.
Secondary sex characteristics, vitellogenin, and plasma sex
steroids are more predictive of an endocrine MOA, although
the high variability seen in plasma sex steroids makes them less
robust and a lower priority than vitellogenin when blood
plasma quantities are limited. In contrast, fecundity is an
indicator of general reproductive condition and not endocrine
specific. Thus, evaluations in the fish reproduction screen are
a combination of endocrine-specific endpoints and apical
endpoints that can be altered by non-endocrine MOAs. With
these apical endpoints and the variability in some of the
observations, weight of the evidence will be employed in
determining the potential of the compound to interact with the
endocrine system of fathead minnow.
Both environmental screens must manage issues related to
the physicochemical properties of various test materials, the
need to develop a solvent-free test system (if possible),
development of analytical methods for dose verification, and
the need to generate range-finding data to assist in dose
selection. Appropriate dose setting is critical to differentiate
systemic toxicity from genuine endocrine-mediated effects.
In 2009, the U.S. EPA released its first list of priority
compounds for endocrine screening. The initial priority list
ENDOCRINE-MEDIATED TOXICITY
consisted of 67 chemicals, primarily pesticide-active ingredients (58) and some pesticide inerts (9). For pesticide-active
ingredients, substantial hazard data are available because of
registration requirements. As part of the EPA EDSP program,
companies were allowed to submit other scientifically relevant
information (OSRI), describing available toxicity data for
a compound and known health risks with an emphasis on
endocrine-related effects. OSRI information for many of the
priority compounds is currently under review by the EPA;
however, it is not clear what information will be viewed as
‘‘functionally equivalent’’ to the Tier I assays. If functionally
equivalent data are available for a compound, companies may
be allowed to omit one or more of the Tier I EDSP assays
based on these existing data.
The goal of the Tier I EDSP is to identify substances that have
the potential to interact with the estrogen, androgen, or thyroid
hormone system for further testing in Tier II. Given that there are
apical endpoints in Tier I assays that may be altered by nonspecific
effects and there is some redundancy to detect endocrine MOAs
across different Tier I assays, it is beneficial to use a weight-ofevidence approach to determine whether there is sufficient data to
warrant Tier II testing (Fig. 1). Previous toxicity data generated
with standardized test guidelines and/or validated methods should
be included in this weight-of-evidence analysis.
One example of applying weight of evidence can be
illustrated with methoxychlor, an estrogenic compound
included in validation studies for several of the Tier I EDSP
screening assays. It is recognized that methoxychlor is
metabolized in vivo to an estrogenic metabolite, 2,2-bis
(p-hydroxyphenyl)-1,1,1-trichloroethane (HPTE). Thus, methoxychlor was not included as part of the estrogen binding
assay validation work; however, other binding assays using
competitive binding to uterine cytosol ERs were negative for
methoxychlor (e.g., Shelby et al., 1996). Methoxychlor was
included in the ER transactivation assay validation, wherein its
estimated EC50 was 1 3 105M. The authors concluded that
methoxychlor exhibited a response curve that was consistent
with other weak estrogens. Methoxychlor was positive in some
AR binding assays (Charles et al., 2005), but it was negative in
the aromatase assay. In vivo, methoxychlor was positive in the
uterotrophic assay (Kanno et al., 2003a,b), although it was
more readily detected with oral exposure than sc injection and
sc injection is the preferred route for administration in the EPA
uterotrophic test guideline. Methoxychlor was negative in the
Hershberger assay at doses that caused a significant increase in
liver weights (Charles et al., 2005) and positive in both the
male and the female pubertal assays. The estrogenic properties
of methoxychlor also were identified in the fish reproduction assay. There were no data on methoxychlor’s effects in
the steroidogenesis or amphibian metamorphosis assays. Based
on a weight-of-evidence approach, methoxychlor would be
identified as an estrogenic compound with possibly a mixed
mode of action that affects androgen-dependent endpoints.
Except where indicated by other citations, results of these
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FIG. 1. This diagram illustrates some of the redundancy in Tier I EDSP
assays. In a weight-of-evidence approach, a pattern of effects across related
assays is easily interpreted (e.g., positive ER binding, ER transactivation, and
uterotrophic assays coupled with estrogenic responses in the pubertal and fish
short-term reproduction assays). However, in many cases, the pattern will not
be as easily discerned. If results of in vitro and in vivo assays disagree, weight
of evidence requires careful consideration of the in vivo results and previous
toxicity data when determining whether the compound is a potential EAC and
warrants Tier II testing. Aside from the expected MOA relationships depicted
in this figure, other, more complex MOA relationships are not shown (e.g., the
ability to detect agents with mixed MOAs like methoxychlor or HPG axis
compounds like atrazine).
validation studies can be located online (http://www.epa.gov/
scipoly/oscpendo/).
Based on a weight-of-evidence approach, compounds
deemed positive in Tier I will undergo Tier II testing to
characterize hazards and develop data for risk assessment. The
specific tests that will be included in Tier II have not been
definitively identified. Possible tests under consideration for
inclusion in Tier II include the two-generation mammalian (rat)
reproductive toxicity study, the two-generation amphibian
(frog) reproductive toxicity study, the two-generation avian
reproductive toxicity study, fish life cycle toxicity study, and
the invertebrate (mysid) life cycle toxicity study.
LOW-DOSE ISSUE AND MIXTURES
Concern over endocrine disruptors has been further
heightened because of reports that such compounds can elicit
unique low-dose effects and exhibit nonmonotonic doseresponse relationships that may not be detected in regulatory
toxicity studies, which are typically conducted at higher dose
levels. The low-dose hypothesis has been a contentious issue
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MARTY, CARNEY, AND ROWLANDS
among some toxicologists. The National Toxicology Program
(NTP), with support from the U.S. EPA, organized a panel to
evaluate the scientific evidence on the low-dose effects of
endocrine disruptors (Melnick et al., 2002), defining low-dose
effects as ‘‘biologic changes that occur in the range of human
exposures or at doses lower than those typically used in the
standard testing paradigm of the U.S. EPA for evaluating
reproductive and developmental toxicity.’’ The panel
concluded that (1) there was evidence for low-dose effects in
laboratory animals, although in some cases, these effects have
not been replicated; (2) the shape of the dose-response curve
depended on the endpoint and dosing regimen used; and (3) the
current reproductive/developmental toxicity testing paradigm
should be reviewed and possibly revised with respect to dose
selection, animal models used, ages evaluated, and endpoints
measured. Some reasons identified for the inability to
reproduce low-dose effects across studies included intrauterine
position, strain and substrain differences in response, dietary
levels of phytoestrogens or differences in caloric intake,
differences in housing conditions (caging, bedding, etc.), and
seasonal variation. The statistics subgroup (Haseman et al.,
2001) also identified a number of important factors that should
be considered in the design, analysis, and interpretation of
experimental studies to help resolve low-dose issues, including
study sensitivity (power), reproducibility, control for litter
effect, quality control in data collection and management,
procedures for data analysis, interpretation of the significance
of biological changes, and data selectivity. Ultimately, the U.S.
EPA concluded, ‘‘Until there is an improved scientific
understanding of the low-dose hypothesis, EPA believes that
it would be premature to require routine testing of substances
for low-dose effects in the Endocrine Disruptor Screening
Program’’ (U.S. EPA, 2002).
Following the NTP low-dose report, a number of key
studies by Naciff et al. (2002, 2009, 2010) have examined
changes in expression of up to 38,500 genes in various cell
lines or uterine tissues exposed to very low doses of EACs,
including BPA at concentrations down to 1nM. These studies
collectively provide support for monotonic dose-responses for
these compounds and showed negligible changes in gene
expression at very low-dose levels. Furthermore, several
comprehensive weight-of-evidence reviews for BPA have
concluded that nonmonotonic dose-response curves and lowdose effects from BPA are not well supported by available
data (Goodman et al., 2009; Kamrin, 2004, 2007). The Center
for the Evaluation of Risks to Human Reproduction (Chapin
et al., 2008) cited many of the issues related to using low-dose
study data when evaluating BPA. Despite these reports, the
Endocrine Society issued a scientific statement on endocrinedisrupting chemicals that included support of the low-dose
hypothesis (Diamanti-Kandarakis et al., 2009); thus, the
debate continues.
The low-dose controversy has evolved to consider issues
related to exposures to mixtures of EACs at low, environmentally
relevant dose levels. Although risk assessment approaches for
cumulative and aggregate risk have been in place for many
years (reviewed in Lipscomb et al., 2010), interest in
cumulative exposures to multiple EACs has received increased
attention in recent years (Carney et al., 2010; Hotchkiss et al.,
2008; Kortenkamp, 2008). Mixtures have been reported to
have additive effects (e.g., Gennings et al., 2004), less than
additive effects (e.g., Charles et al., 2007), greater than additive
effects (e.g., Laetz et al., 2009), or even counteracting effects
(e.g., Tanida et al., 2009). Overall, data indicate that the
interactions of chemicals in mixtures are dose dependent.
Using mixture data in the context of environmental risk
assessment is even more complex when one considers natural
exposures to EACs in food and pharmaceuticals (Gaido et al.,
1998; Safe et al., 1997). Exposures to environmental chemicals
typically involve low doses of each chemical within a mixture
and, given the safety factors used in risk assessment, exposures
generally below the threshold for toxicity. Thus, mixtures
research requires a focus on the behavior of mixtures at
environmentally relevant exposures because results obtained at
high-dose levels may not apply in real-world scenarios. This
concept also was espoused by the Expert Working Group on
Mixtures Toxicology, which was convened by the Society of
Toxicology and the Society for Environmental Toxicology and
Chemistry (Teuschler et al., 2002).
HISTORICAL DEVELOPMENTS IN THE BASIC SCIENCE
OF ENDOCRINE SIGNALING
When the history of endocrinology is written, it is likely that
the 20th century will be neatly divided into two epochs with the
first half being the era of biochemistry when the hormones were
discovered and the emphasis was on proteins and the second half
of the 20th century, the era of genomics when the genes for the
hormone and nuclear receptors (NRs) were identified. The word
hormone was first coined in 1905, but it was not until 1926 that
the first hormones were characterized in detail when Kendeall
and Reichstein isolated and determined the structures of
cortisone and thyroxine (The Nobel Assembly at Karolinska
Institutet, 1950). Soon thereafter, in 1929, Adolf Butenandt and
Edward Adelbert Doisy independently isolated and determined
the structure of estrogen (Tata, 2005). These discoveries along
with other experimental observations allowed Clark to propose
his model for receptor-mediated processes in his publication of
the theory of occupancy. The receptor theory has been refined
since the 1930’s along with the vocabulary used to classify
chemicals that bind to NRs or ligands, which are now classified
into agonist, antagonist, partial agonist, and inverse agonist
based on the biological effects elicited through NRs (Table 2).
Potential EAC’s may produce their effects through interactions with NRs. Under normal physiological conditions, the
intracellular NRs are bound by their endogenous signaling
molecules, or receptor ligands, such as the steroid hormones,
ENDOCRINE-MEDIATED TOXICITY
TABLE 2
Properties of Chemicals that Affect NRs
Ligand
Affinity
Agonists
Antagonists
Partial agonists
Inverse agonists
Efficacy
Potency
Chemical that binds reversibly to an NR
A ligand’s capacity to bind to an NR
Ligand that produce an NR conformation
that is transcriptionally active
Ligands that prevent agonist binding or induce
an NR conformation resulting in an action
distinct from that produced by an agonist
Ligands that possess weak agonist activity
Ligands that induce an NR conformation that
results in transcriptional repression
A measure of the capacity of the ligand to
activate, prevent activation, or inactivate an NR
A measure of the concentration or amount of
ligand needed to produce an effect through an NR
thyroid hormones, retinoids, vitamin D, and other as yet
identified endogenous ligands. These receptor ligands either
diffuse directly across the plasma membrane of cells or are
sometimes produced in the same cells and bind to the NR. The
commonly understood mechanism of action for NRs involves
the initial ligand binding that results in receptor activation,
binding at specific DNA response elements in nuclear
chromatin, followed by recruitment of a variety of cofactors
that modify the chromosomal organization and interaction with
the transcription complex, thereby altering the rates of
transcription of genes. So, binding of a ligand to an NR is
only the first step in a complex cascade of events leading to
changes in gene expression and subsequent cell signaling. With
the completion of the human genome, it is now clear that there
are some 48 NRs expressed in humans, of which only half have
identified ligands such as steroid hormone receptors that
constitute only a minority of this large superfamily of receptors
(Gronemeyer et al., 2004).
Like the NR superfamily, coregulators are themselves a large
and mixed family of proteins and possess a varied array of
enzymatic activities necessary for chromosomal maintenance
and gene regulation (Roeder, 2005). They can be categorized
based upon their wide-ranging functional properties including
acetyltransferases, ubiquitin ligases, ATP-coupled chromatin
remodeling complexes, protein methylases, cell cycle regulators, RNA helicases, and bridging proteins that facilitate
direct contact with components of the basal transcription
machinery (McKenna and O’Malley, 2003). There are
currently 285 coregulators identified and many are expressed
differentially by cell type, tissue type, gender, developmental
stage, and species (Lonard et al., 2007; Smith and O’Malley,
2004). Moreover, coregulators can be differentially recruited to
the same receptor by ligand-induced conformational changes in
the NR. Thus, with 48 NRs and 285 plus coregulators, the
complexity of these endocrine signaling pathways is apparent.
It is important to note that all these concepts of agonists,
antagonists, partial agonists/antagonists, or inverse agonists are
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conditional and specific to a given gene, in a specific cell and
tissue under specific conditions. For example, a given ligand
may be an agonist for a gene in a cell and at the same time be
an antagonist for a different gene in the same cell; or a ligand
can be an agonist for a gene in one cell type and an antagonist
for the same gene in a different cell type or even the same cell
type but under different conditions. Potency is a conditional
measure that depends on parameters of affinity and efficacy, as
well as tissue receptor numbers and ligand bioavailability.
Thus, the intrinsic efficacy of an EAC, whether agonist,
antagonist, partial agonist/antagonist, or inverse agonist for
a given NR, is dependent upon both chemical properties of the
EAC and the biological environment of the NR (Jordan, 2003).
These differences in the properties of ligands are often lost in
the discussion of EACs. An EAC’s potency is dependent on its
affinity for an NR, and some putative EACs have ER binding
affinities that are 10,000–100,000 lower than ER binding
affinities of potent, natural, or contraceptive estrogens
(Gutendorf and Westendorf, 2001). For example, the affinity
of BPA for ERs is approximately 10,000-fold weaker than that
of estradiol (Kuiper et al., 1998). What this means in practical
terms is that a cell would need to be exposed to 10,000
molecules of BPA to produce the same effect through the ER
that one molecule of estradiol would produce through that same
receptor. Furthermore, the amount of BPA that one would need
to consume, breath, or absorb through the skin would be even
greater because the absorption, distribution metabolism, and
elimination of the BPA would limit the amount of BPA
available to the cells. Even if an EAC was highly potent in one
cell type, it does not mean it will have the same efficacy in
another cell type or the same cell type under different
conditions as efficacy is independent of ligand affinity and
instead reflects a ligand’s ability to induce a stable coactivator
binding site on the NR (discussed further below). Thus, two
ligands can have similar potencies for inducing a response
through an NR, but the efficacy for these two ligands inducing
the response can vary greatly and this can have a profound
impact on the ability of these ligands to produce the same
apical effects through the NR. The lower potencies and
different efficacies of EACs relative to the natural hormones
need to be understood and considered in the discussions of the
potential impact of EACs on human health and the environment.
It took another three decades after the elucidation of the
structure of estrogen before the ER was isolated in 1961 by
Elwood Jensen (Tata, 2005) and another two decades passed
before the new tools of molecular biology allowed Pierre
Chambon’s group to clone the gene for the human ER (Green
et al., 1986). The 1980’s ushered in an era of rapid discovery of
the receptor genes for most of the major hormones, which was
followed in the 1990’s by the cloning of novel nuclear orphan
receptors for which there are no known ligands. It was also
during this period, in 1996, that Jan-Åke Gustaffson’s group
cloned a second ER, now called ERb (Kuiper et al., 1996),
requiring the observations of the previous decades of estrogen
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studies to be viewed through a new lens. No doubt there will be
additional chemicals identified as ligands for some of the 48
NRs identified in humans, thereby necessitating an understanding of their potential for behaving as EACs.
ESTROGENS: FACTORS AFFECTING AFFINITY, POTENCY,
AND EFFICACY
One of the best understood examples of an endocrine
signaling system that brings all these concepts together is that
of estrogens. A half-century has passed since the initial reports
of a cognate ER, and it is now known that ERa and ERb
exist, as well as less well-understood additional subtypes. The
distribution of these two receptors is variable, with tissues such
as bone having predominantly the b isoform, uterus containing
mainly a, but many tissues expressing both a and b.
Furthermore, a and b often regulate opposing functions
providing a feedback loop for finite control of estrogen
signaling. The biological effects elicited by specific ER ligands
are highly cell and tissue specific. One of the best known
examples of this phenomenon is the drug tamoxifen, which has
been used for treatment of early stage breast cancer and as
a chemoprevention agent for women at high risk for this
disease (Jordan, 2003). Tamoxifen is an ER antagonist in breast
cancer but an ER agonist in the bone and uterus (Fong et al.,
2007). Another example is the phytoestrogen resveratrol,
which shows more activity in cells that uniquely express
ERb or that express higher levels of ERb than ERa and also
appears to exhibit estrogen antagonist activity for ERa in
certain promoter contexts (Bowers et al., 2000). In addition, only
a small percentage of ER ligands are pure estrogen agonists or
antagonists. Most phytoestrogens and xenoestrogens are partial
agonists, meaning that they are unable to induce a maximal
response matching that of the endogenous ligand, even at very
high concentrations (Zhu, 2005). The latter compounds tend to
act as agonists under some conditions but act as antagonists
when their concentration is very high relative to more potent
ligands. The research from numerous laboratories over the last
five decades has revealed the explanations for many of these
observations.
Differences in the tissue-specific receptor agonist and
antagonist activities have long been recognized for the steroid
hormone receptors and form the basis for the development of
drugs for treatment of hormone-dependent diseases (Ai et al.,
2009; Nettles and Greene, 2005). Tamoxifen’s tissue-specific
activity is now thought to be because of recruitment of different
coregulators to the ER by tamoxifen in breast tissue versus
uterine and bone tissue (Gronemeyer et al., 2004). Tamoxifen
was the first of what are now called selective ER modulators or
SERMs where the cellular response to SERMs is determined
by cell type– and promoter-specific differences in coregulator
recruitment by the ER. Thus, coregulators extend the actions of
the same hormone receptor by producing cell type–specific
FIG. 2. The actions of nuclear receptors are ligand dependent and cell
context dependent. Nuclear receptors (R) form a different tertiary structure
depending on the ligand that activates them (EAC-1 or EAC-2). This in turn
causes the receptors to recruit different coregulator complexes, e.g., coactivator
complexes or corepressor complexes that lead to gene induction or repression,
respectively. In Cell A, the two different ligands (EAC-1 and EAC-2) bind to
the same receptor but induce two different tertiary structures, which in turn
cause the receptor to recruit a coactivator complex in the case of EAC-1 bound
receptor or a corepressor complex in the case of EAC-2 bound receptor.
Whereas in Cell B, the opposite occurs with EAC-1 producing a corepressor
complex recruitment through the receptor, but EAC-2 produces a coactivator
complex recruitment through the receptor. Thus, an individual EAC can act as
an agonist, antagonist, partial agonist/antagonist, or inverse agonist depending
on the coregulators present in the specific cell. Although not shown, the gene
architecture influences nuclear receptor-coregulator recruitment as well.
Coactivator (coactivator complex); corepressor (corepressor complex); TBP,
A, B, F, pol II (RNA polymerase II holoenzyme: RNA polymerase II and
general transcription factors); TATA (TATA box promoter site); transcription
(RNA transcript).
positive or negative stimuli from the same hormone ligand or
different receptor ligands depending on the cell-specific
coregulator availability and/or ligand-specific induced receptor
conformations with differential coregulator affinities (Fig. 2).
Selective receptor modulators (SRMs) have been observed for
other NRs including peroxisome proliferator–activated receptors, ARs, progesterone receptors, as well as the AhR
(Gronemeyer et al., 2004; Zhang et al., 2008). Thus, SRM
appears to be a common feature of ligand-activated transcription factors that explains the observed selectivity that different
ligands confer to NRs on modulation of gene expression and
the resultant tissue changes both within and across tissue and
cell types (Ai et al., 2009; Zhang et al., 2008). No doubt, EACs
will share similar characteristics with SRMs.
All the discoveries discussed above and more have been
harnessed to develop in vitro and in vivo assays to identify
chemicals with endocrine-modulating activities. The very
ENDOCRINE-MEDIATED TOXICITY
nature of an NR being a trans-acting transcriptional regulator
allows for their exogenous expression in numerous cell types
and cell lines to study the effects of potential ligands as
agonists, antagonists, etc. These studies have led to the
adoption of some assays into the EDSP battery of tests and
many others that are used in the ToxCast panel of assays. The
cloning of these receptors has also facilitated the development
of NR knockout (KO) mouse models that provided important
confirmatory evidence for the physiological role of hormones
in development, reproduction, and normal physiology as well
as some unexpected observations.
With the advent of the KO models, it was possible to
develop transgenic receptor knock in mouse models that
express human NRs in place of their endogenous mouse
receptor. The development of the ER KO mouse (ERKO
mouse) by the National Institute of Environmental Health
Sciences group led by Ken Korach demonstrated that both
sexes of these mutant animals are infertile and show a variety
of phenotypic changes, some of which are associated with the
gonads, mammary glands, reproductive tracts, and skeletal
tissues (Lubahn et al., 1993). The ERKO mice and other
hormone receptor KO mouse models can be used to confirm
the endocrine effects of putative EACs and demonstrate
important differences in the effects produced by endocrine
chemicals compared with endogenous hormones. By knocking
in the human gene for an NR into the appropriate receptor KO
mouse, new transgenic mouse models are made that allow
comparison of putative EACs acting through the human
receptor or mouse receptor. The observations from such
studies are important to consider in the weight of evidence
for predicting the effects of EACs to humans. Future mouse
models are being developed that express the human proteins for
multiple NRs and accessory proteins, and these will allow
further refinement of the effects of EACs on human pathways.
In addition, the technology is now available to make KO rats
and other species, and this will be followed by the ability to
humanize these models as well, thereby furthering our ability to
understand the human relevance for EACs identified in wildtype laboratory models and in vitro assays.
FUTURE DIRECTIONS IN ENDOCRINE DISRUPTOR
SCREENING AND TESTING
Endocrinology has come far since the term hormones were
first introduced in 1905. What started as a simple concept has
evolved into a very complex array of hormones, receptors, and
cofactors. Because a chemical ligand for an NR can have
multiple intrinsic efficacies (i.e., induce several NR responses)
depending upon the cellular and tissue environment of the NR,
relying upon in vitro cell and biochemical assays to define EACs
has important implications. With suspect EACs possessing the
potential for multiple induced NR responses, the validity for
transferring inferences about the efficacy of the EAC-induced
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effect obtained for one response to another response is
questionable. Moreover, the problem of confounding is even
more problematic when considering that measured parameters of
in vitro high-throughput screening systems are by nature different from those found for receptor systems in vivo. Therefore,
the results from these in vitro screening systems may seldom
accurately reflect the intrinsic efficacies of NR ligands in vivo
and challenge the validity for extrapolating observations to
humans. Even if a single in vitro system was reflective of
a human response under certain conditions, the fact that the
intrinsic efficacy of an EAC is dependent upon the cellular
context of the NR means that multiple systems and or conditions
would need to be employed to ensure that all the possible
conditions in vitro reflect similar conditions in humans in vivo.
Although this review only touched briefly on some of the
key developments to date concerning endocrine-mediated
toxicity and safety assessment, it should be readily apparent
that the issue has raised more questions and challenges than
have been answered. Most likely, we are only at the very
beginning of a long journey toward improved understanding of
the basic mechanisms of endocrine-mediated toxicity and the
extent to which relevant exposures to hormonally active agents
impact human and environmental health. As we forge ahead,
this active and often controversial topic has already accelerated
the science of toxicology and risk assessment and has
redirected it into some fundamentally different directions.
One such direction is down. This is not meant in a negative
sense but refers to the examination of much lower, environmentally relevant dose levels than have been examined in the
traditional high-dose toxicology paradigm. This new direction
was spurred by research, albeit controversial, that hormonally
active agents might not adhere to the monotonic dose-response
behavior, which has been a fundamental principle of toxicology
since its inception as a field of science. However, further progress in this area is possible thanks to modern technology, such
as toxicogenomics, which allows one to examine for precursor
changes on the cellular and molecular level (Naciff et al., 2009)
and thus help to better understand the fundamental biology
underlying dose-response for environmentally relevant exposures. In addition, the increased utilization of human, rather than
animal, cells holds promise for enhancing human relevance
even further.
Another research direction points upward, namely, toward
the investigation of higher numbers of chemicals and other
stressors and their potential cumulative effects. For years,
toxicologists have tended to avoid studying cumulative effects
because it was considered an intractable problem. This was
especially so for hormonally active agents because of the vast
array of ligands and multiplicity of sources, which includes not
only man-made chemicals but also compounds in plants,
endogenous hormones, and even nonchemical stressors.
However, because of new high-throughput tools, refined
statistical methods, and economical experimental designs for
mixture studies (Gennings, 1996), questions about combined
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exposures to hormonally active agents can be addressed in
a scientifically rigorous and efficient manner.
This movement toward evaluating more chemicals is also
being combined with a fundamental change toward mechanismbased testing (inside out) versus the traditional adverse effectsbased system (outside in) as described in the landmark report
by the U.S. NRC Committee on Toxicity Testing and 1365
Assessment of Environmental Agents, Board of Environmental
Studies and Toxicology, Institute for Laboratory Animal
Research (2007) report entitled ‘‘Toxicity Testing in the 21st
Century: A Vision and a Strategy.’’ Within this report,
endocrine screening is highlighted as an area of toxicological
concern that could benefit from the use of molecular-based
screening methods. Progress has been made toward implementing
this approach. EPA’s ToxCast program contains numerous
endocrine-related screening assays, including ER and AR
receptor binding assays, ER, AR, and thyroid receptor reporter
gene assays, an assay for thyrotropin-releasing hormone
receptor, and an aromatase enzymatic activity assay (Houck,
2009). These assays are conducted in high-throughput mode,
so results are generated quickly. However, there are concerns
about relying on these in vitro and alternative assays for
regulatory decisions in determining potential endocrine activity
(e.g., Bus and Becker, 2009) because of the lack of
toxicokinetic considerations (absorption, distribution, metabolism,
and excretion), the imperfect concordance of ToxCast assays
with in vivo data, and achieving agreement on a mechanismbased definition of adverse. In the short run, these data may
contribute to the weight of evidence for some test compounds
in determining potential endocrine activity. Furthermore, if the
vision of ‘‘Toxicology in the 21st Century’’ is realized, such an
approach could prove to be far more efficient and cost effective
than the existing EDSP assays, some of which are resource
intensive and use large numbers of animals to screen for
a limited number of MOAs.
Finally, future research in this area needs to look outward,
meaning that information on specific components of the
endocrine system (e.g., specific cell culture responses) needs
to be interpreted in the context of whole-body physiology,
preferably that of humans. In the case of environmental
species effects, integrative frameworks such as the ‘‘adverse
outcomes pathway’’ concept have recently been developed to
address links between events at the molecular level all the
way up through adverse outcomes at the population level
(Ankley et al., 2010). Because the endocrine system
integrates responses across distant cells, tissues, and organs,
a reductionist approach that solely focuses on isolated
components of an endocrine circuit can generate incomplete
answers and may even guide public health professionals in the
wrong direction. In this same vein, basic principles of
toxicology and in particular, toxicokinetics, should not be
forgotten and need to be part of any research program if it is
to be meaningful for risk assessment. For example, recent
research on the toxicokinetics of BPA and genistein has
highlighted the tremendous impact of hepatic first pass
metabolism in converting these compounds to their
biologically inactive glucuronide forms (Teeguarden and
Barton, 2004; Teeguarden et al., 2005). Therefore, route of
exposure can have a dramatic effect on the percentage of free,
biologically active compound and consequently the effects
produced. The key point is that as we delve deeper into the
details of individual mechanisms, it is essential that these new
findings be interpreted in a holistic, physiological framework.
Fortunately, new tools in exposure modeling, physiologically
based pharmacokinetic modeling, and systems biology are
increasingly available to assist with experimental design, data
integration, and data interpretation to achieve this goal.
CONCLUSION
Concerns about EACs have come to the forefront of
toxicology over the past 15 years. The precise impact of
EACs is difficult to discern as EACs can be beneficial or
adverse, producing different effects in different tissues at
different life stages. To complicate this issue further, exposure
to low-level environmental EACs occurs in an environment in
which natural EACs (e.g., phytoestrogens) and/or exposures
to pharmaceuticals contribute to the background hormonal
mileau. The focus on endocrine disruption represents a new
paradigm in toxicology, wherein the MOA, rather than
toxicological outcome, is the focus of chemical screening.
Development of the EDSP has been more complex than
originally envisioned, and the difficulty in identifying assays,
efficiently conducting those assays, and data interpretation
have become apparent. As endocrine issues have increased in
scope, increased research has led to a better understanding of
receptor biology. To take full advantage of this new
knowledge and efficiently meet the needs to screen compounds for endocrine activity, the EDSP should be iterative
and allow for reevaluation of earlier assumptions based on the
state of the science. Clearly, the past 15 years have
demonstrated that legislative mandates for specific types of
toxicity testing that precede scientific understanding are
difficult to implement and limit the latitude to adapt to
evolving science. Even so, the significant resources dedicated
to developing and validating the panel of EDSP assays have
led to an improved understanding of the challenges involved
with implementing and interpreting EDSP assays, and these
lessons will be beneficial to eventual implementation of the
EPA’s ToxCast assays and human primary cell assays as
proposed in the National Research Council report Toxicity
Testing in the 21st Century.
ACKNOWLEDGMENTS
The authors would like to thank Drs James Bus, Matt
Lebaron, and Katie Coady for their thoughtful reviews of the
manuscript and Dr David Geter for contributing Figure 1.
ENDOCRINE-MEDIATED TOXICITY
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