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M. R. Preston
DDT family) (Botham et al. 1999; Vos et al. 2000). This was introduced to a lake which,
despite being already contaminated by agricultural runoff and sewage effluent, had a
successful alligator population. By 1984 the juvenile alligator population had declined
by around 90%. Eggs collected from the lake region had high contaminant concentrations, and abnormalities were also noted in the embryos and young. Males from Lake
Apopka eggs had poorly organized testes with aberrant structures while six-monthold females had abnormal ovaries with prominent polyovular follicles and an unusually large number of multinucleated oocytes (Botham et al. 1999). Older animals also
showed abnormalities that could be linked to endocrine disruption (Semenza et al.
1997). For example, concentrations of plasma testosterone were higher in females and
lower in males than in cleaner environments, whilst 17/3-oestradiol were higher than
normal in males. Males also showed a link between reduced penis length and androgen doses (Botham et al. 1999).
Less research has been performed on the influence of endocrine-disrupting chemicals and amphibians, though there is increasing concern about world-wide declines
in amphibian populations (see e.g. Carey 2000). This is in part because of the complex sex determination factors in these animals that include temperature as a major
parameter. Nevertheless, links have been made between the presence of endocrinedisrupting chemicals and various abnormalities in frogs (Xenopus laevis, Palmer and
Palmer 1995; Acris crepitans, Reeder et al. 1998) and salamanders (Clark et al. 1998).
Pickford and Morris (1999) have hypothesized that endocrine-disrupting chemicals
may disrupt progesterone-induced oocyte maturation in the adult amphibian ovary
and have tested this with the African clawed frog (Xenopus laevis) and the pro-oestrogenic pesticide methoxychlor. Effects were noted at very low concentrations (mean
inhibitive concentration 72 nM), and the effects of the chemical were shown to be dosedependent, reversible and early acting.
Recently a number of researchers (e.g. Lutz and Kloas 1999; Kloas et a1.1999; Palmer
et al. 1998) have examined the use of amphibians as a model to study endocrine-disrupting chemicals. Such models are based around the mechanisms of vitellogenin
production, binding to liver receptors and in vivo effects on sexual development caused
by larval exposure to chemicals. An overview of a workshop for detecting potential
(anti- )oestrogenic/androgenic chemicals in wildlife has also recently been provided
by Ankley et al. (1998).
13.4.3
Birds
The damaging effects of organochlorine pesticides (notably DDT/DDE, PCBs and the
cyclodiene pesticides) on bird populations were firmly established after the major
declines in top predator bird populations in many industrialized countries during the
1950S and 1960s (Botham et al.1999). Effects include eggshell thinning (Blus et al. 1997;
Grasman et al.1998; Vos et al. 2000), female-female pairing (Hunt and Hunt 1977) and
supernormal clutches (Fry and Toone 1981; Fryet al. 1987; Dawson 2000). Species such
as peregrine falcons (Falco perigrinus), cormorants (e.g. Phalacrocorax auritus) and
brown pelicans (Pelecanus occidentalis) have been shown to be particularly vulnerable though a variety of other species (e.g. dippers, Cinclus cinclus; Ormerod et al. 2000)
may also be at risk. Whether such effects are due to endocrine-disruption mechanisms
M. R. Preston
DDT family) (Botham et al. 1999; Vos et al. 2000). This was introduced to a lake which,
despite being already contaminated by agricultural runoff and sewage effluent, had a
successful alligator population. By 1984 the juvenile alligator population had declined
by around 90%. Eggs collected from the lake region had high contaminant concentrations, and abnormalities were also noted in the embryos and young. Males from Lake
Apopka eggs had poorly organized testes with aberrant structures while six-monthold females had abnormal ovaries with prominent polyovular follicles and an unusually large number of multinucleated oocytes (Botham et al. 1999). Older animals also
showed abnormalities that could be linked to endocrine disruption (Semenza et al.
1997). For example, concentrations of plasma testosterone were higher in females and
lower in males than in cleaner environments, whilst 17/3-oestradiol were higher than
normal in males. Males also showed a link between reduced penis length and androgen doses (Botham et al. 1999).
Less research has been performed on the influence of endocrine-disrupting chemicals and amphibians, though there is increasing concern about world-wide declines
in amphibian populations (see e.g. Carey 2000). This is in part because of the complex sex determination factors in these animals that include temperature as a major
parameter. Nevertheless, links have been made between the presence of endocrinedisrupting chemicals and various abnormalities in frogs (Xenopus laevis, Palmer and
Palmer 1995; Acris crepitans, Reeder et al. 1998) and salamanders (Clark et al. 1998).
Pickford and Morris (1999) have hypothesized that endocrine-disrupting chemicals
may disrupt progesterone-induced oocyte maturation in the adult amphibian ovary
and have tested this with the African clawed frog (Xenopus laevis) and the pro-oestrogenic pesticide methoxychlor. Effects were noted at very low concentrations (mean
inhibitive concentration 72 nM), and the effects of the chemical were shown to be dosedependent, reversible and early acting.
Recently a number of researchers (e.g. Lutz and Kloas 1999; Kloas et a1.1999; Palmer
et al. 1998) have examined the use of amphibians as a model to study endocrine-disrupting chemicals. Such models are based around the mechanisms of vitellogenin
production, binding to liver receptors and in vivo effects on sexual development caused
by larval exposure to chemicals. An overview of a workshop for detecting potential
(anti- )oestrogenic/androgenic chemicals in wildlife has also recently been provided
by Ankley et al. (1998).
13.4.3
Birds
The damaging effects of organochlorine pesticides (notably DDT/DDE, PCBs and the
cyclodiene pesticides) on bird populations were firmly established after the major
declines in top predator bird populations in many industrialized countries during the
1950S and 1960s (Botham et al.1999). Effects include eggshell thinning (Blus et al. 1997;
Grasman et al.1998; Vos et al. 2000), female-female pairing (Hunt and Hunt 1977) and
supernormal clutches (Fry and Toone 1981; Fryet al. 1987; Dawson 2000). Species such
as peregrine falcons (Falco perigrinus), cormorants (e.g. Phalacrocorax auritus) and
brown pelicans (Pelecanus occidentalis) have been shown to be particularly vulnerable though a variety of other species (e.g. dippers, Cinclus cinclus; Ormerod et al. 2000)
may also be at risk. Whether such effects are due to endocrine-disruption mechanisms