Endocrine-Disrupting Chemicals and Their Environmental Impacts - Organic Pollutants: An Ecotoxicological Perspective - page 267

by blocking the activation of the estrogen receptor or by binding the aryl hydro-
carbon (Ah) receptor, in turn leading to induction of Ah-responsive genes that can
have a spectrum of antiestrogenic effects (Lerner et al. 1958, in Wakeling 2000).
Antiestrogens create an androgenic environment, producing symptoms similar to
those of androgenic exposure. Antiestrogenic chemicals known to enter the environ-
ment include pharmaceuticals, such as tamoxifen and fulvestrant, used to treat breast
cancer; raloxifene, which is used in the prevention of osteoporosis; and some of the
polyaromatic hydrocarbons (PAH) such as anthracene (Tran et al. 1996).
Chemicals with antiandrogenic activity include pharmaceuticals developed
as anticancer agents (e.g., flutamide, Neri and Monahan 1972; Neri et al. 1972, in
Lutsky et al . 1975) and 179-methyltestosterone used to treat testosterone deficiency
(Katsiadaki et al. 2006). Other antiandrogens include various pesticides such as the
p,p ′ - DDE metabolite of DDT, the herbicides linuron and diuron, and metabolites of
the fungicide vinclozolin (Gray et al . 1994). Antiandrogens create a similar over-
all effect to estrogens (Kelce et al . 1995), and it been hypothesized that some of
the feminized effects seen in wildlife populations may result from chemicals block-
ing the androgen receptor rather than as a consequence of exposure to (or possibly
in addition to) environmental estrogens (Sohoni and Sumpter 1998; Jobling et al.,
submitted). An extensive study on wastewater treatment works (WWTW) effluents
in the United Kingdom has found very widespread antiandrogenic activity in these
discharges (Johnson et al. 2004; see case example for the feminization of fish later in
chapter). There has also been increasing evidence to support links between increases
in the group of disorders referred to as testicular dysgenesis syndrome (TDS) in
humans, which originate during fetal life, and exposure to environmental chemicals
with antiandrogenic activity (Fisch and Golden 2003; Sharpe and Skakkebaek 2003;
Sharpe and Irvine 2004; Giwercman et al . 2007).
Few environmental androgens have been identified, but one of the best examples
of hormonal disruption in wildlife is an androgenic effect, namely, the induction of
imposex in marine gastropods exposed to the antifouling agent tributyl tin (TBT,
discussed in detail in Section 15.4). Androgenic responses in vertebrate wildlife
are also known to occur, and reported examples include the masculinization of
female mosquito fish, Gambusia affinis holbrooki , living downstream of a paper
mill effluent (Howell et al . 1980), and the masculinization of fathead minnow,
Pimephales promelas , living in waters receiving effluent from cattle feedlots in
the United States (Jegou et al . 2001). In the latter case, the causative chemical was
identified as 17β-trenbolone (TB), a metabolite of trenbolone acetate, an anabolic
steroid used as a growth promoter in beef production (Wilson et al . 2002; Jensen
et al. 2006).
Several groups of chemicals are known that can disrupt thyroid function. Some of
these chemicals have a high degree of structural similarity to thyroid hormones and
act via binding interference with endogenous thyroid hormone receptors. Thyroid
hormones are fundamental in normal development and function of the brain and sex
organs, as well as in metamorphosis in amphibians, and in growth and regulation of
metabolic processes (Brouwer et al . 1998) and, thus, chemicals that interfere with
their functioning can potentially disrupt a very wide range of biological processes.
Developmental effects in wildlife populations indicative of disruptions in the thyroid