Endocrine-Disrupting Chemicals and Their Environmental Impacts - Organic Pollutants: An Ecotoxicological Perspective - page 282

microbial breakdown of APEs and have been known to be estrogenic since 1938
(Dodds and Lawson 1938). As with bisphenol A, NP was rediscovered more recently
as an estrogenic xenobiotic by Soto et al. (1991) in an estrogen-dependent cell prolif-
eration assay. Alkylphenols are relatively weak estrogens, with an affinity for the ER
2,000-100,000-fold less than E 2 (reviewed in Nimrod and Benson 1996). However,
alkyphenols can also interact with the androgen receptor to induce antiandrogenic
effects (Gray et al . 1996). NP induces VTG synthesis in fish at concentrations as low
as 6.1 µg/L for a 14-day exposure and 650 ng/L over for a 3-week exposure (Harries
et al. 2000; Thorpe et al. 2001). NP has also been shown to affect pituitary function
and the release of gonadotrophins (which control the whole reproductive cascade
in vertebrates) in fish at concentrations of only 0.7 μg/L for an 18-week exposure
(Harris et al. 2001). As for many other EDCs, longevity of exposure affects both the
threshold and magnitude of the response; alkylphenols such as NP have been shown
to bioconcentrate in fish up to 34,000-fold (Smith and Hill 2004). In the aquatic
environment, rivers and the sea receive substantial amounts of APEs from WWTWs
and industrial effluent discharges. It is estimated that 60% of the world's production
ends up in the aquatic environment (Uguz et al. 2003). Domestic effluent can contain
up to hundreds of µg APEs/L (Naylor 1995), whereas industrial effluent, especially
that from pulp and textile industries, can contain mg/L concentrations. In some riv-
ers in the United Kingdom that have historically received high-level discharges from
the textile industry, alkylphenolic chemicals were shown to be some of the major
contaminants inducing feminized responses in exposed fish (Harries et al. 1996;
Sheahan et al. 2002).
15.6.8 p h T h a l a T e s
Phthalates are the most abundant synthetic chemicals in the environment (Peakall
1974). Used in lubricating oils, insect repellents, and cosmetics and predominantly
to impart flexibility to plastics, phthalates have been measured in rivers (Sheldon
and Hites 1978; Fatoki and Vernon 1990), drinking waters (Suffet et al . 1980), and
marine environments (Jobling et al . 2002b). Many thousands of tons of plastics
are also disposed of annually in landfill sites, resulting in phthalate esters leach-
ing into and contaminating groundwaters. The estrogenic activity of two phthalate
esters, di- n -butylphthlate (DBP) and butylbenzyl phthalate (BBP), was discovered
by Jobling et al. (1995). Further studies in fish have shown that BBP and diethyl
phthalate (DEP) both induce VTG at an exposure concentration around 100 µg/L via
the water (Harries et al . 2000; Barse et al. 2007). Various in vitro screens and tests
have shown that phthalates mediate their effects via binding to the estrogen receptor
(Jobling et al . 1995; Harries et al. 1997). However, antiestrogenic effects of phtha-
lates also occur in vivo (inhibition of VTG expression). In addition to these estrogenic
effects, some phthalates are also known to be toxic to aquatic organisms (Mayer and
Sanders 1973, in Giam et al . 1978) and mammals (Lee et al. 2007). Recent studies
have shown that DBP, monoethyl phthalate (MEP, a metabolite of DEP), and mono-
(2-ethylhexyl) phthalate (MEHP) can induce DNA damage in human sperm and/
or male rats (Wellejus et al . 2002; Hauser et al. 2007). Environmental concentra-
tions of dimethyl phthalate (DMP), DEP, DBP are reported to be between 0.3 to