Ecotoxicology and Environmental Safety 54 (2003) 302–314
Identification of endocrine-disrupting effects in aquatic vertebrates
and invertebrates: report from the European IDEA project$
H. Segner,a,* K. Caroll,b M. Fenske,a C.R. Janssen,c G. Maack,a D. Pascoe,b C. Sch.afers,d
G.F. Vandenbergh,c M. Watts,b and A. Wenzeld
a
Department of Chemical Ecotoxicology, UFZ Centre for Environmental Research, Permoserstrasse 15, D-04318 Leipzig, Germany
b
School of Biosciences, Cardiff University, P.O. Box 915, Cardiff CF10 3TL, UK
c
Laboratory of Environmental Toxicology and Aquatic Ecology, Ghent University, J. Plateaustraat 22, B-9000 Ghent, Belgium
d
IME Fraunhofer Institute for Molecular Biology and Applied Ecology, Auf dem Aberg 1, D-57392 Schmallenberg, Germany
Received 31 July 2001; received in revised form 24 May 2002; accepted 14 June 2002
Abstract
The EU-funded project IDEA aimed to evaluate (a) what parameters and endpoints allow the detection of endocrine-mediated
developmental and reproductive effects of (xeno)estrogens in life cycle- and life stage-specific toxicity tests with the zebrafish Danio
rerio, a small laboratory fish used in many ecotoxicity test guidelines, and (b) whether substances that act as estrogens in vertebrates
may also adversely affect the development, differentiation, and reproduction of aquatic invertebrates. The invertebrate species
investigated included Hydra vulgaris, Gammarus pulex, Chironomus riparius, Hyalella azteca, and Lymnaea stagnalis. The animals
were exposed to the model estrogenic chemicals ethynylestradiol (EE2), bisphenol A (BPA), and octylphenol (OP), which exert their
endocrine activity in vertebrates through the estrogen receptor. As endpoints, developmental and reproductive parameters at the
organism level as well as molecular and cellular parameters were measured. Life cycle exposure of zebrafish to (xeno)estrogens
induced a specific, partly irreversible response pattern, consisting mainly of (a) induction of vitellogenin (VTG), (b) alterations of
gonad differentiation, (c) delay of first spawning, and (d) reduced fertilization success. The effects of EE2 on zebrafish were
expressed at environmentally realistic concentrations, while BPA and OP became effective at concentrations higher than those
usually found in the environment. The vitellogenic response was equally sensitive as the reproductive parameters in the case of EE2,
but VTG was more sensitive in the case of BPA. Partial life cycle exposure of zebrafish had lasting effects on fish development and
reproduction only when the fish were exposed during the stage of juvenile bisexual gonad differentiation. In (partial) life cyle and
multigeneration studies with invertebrates, (xeno)estrogenic impact was assessed by a range of developmental and reproductive
parameters including hatching, growth, moulting, mating behavior, and egg number. Several parameters were found to be
responsive to (xeno)estrogens; however, most effects were induced only at higher, probably nonphysiological concentrations. Lowdose effects were observed in full life cycle experiments, particularly in the second generation. It remains to be established whether
the estrogen-induced alterations in the invertebrate species indeed do result from disturbances of the endocrine system. The findings
of the present research project support the development of appropriate testing methodologies for substances with estrogenic activity.
r 2002 Elsevier Science (USA). All rights reserved.
1. Introduction
There is considerable concern over the environmental
occurrence of endocrine-disrupting chemicals (EDCs),
$
This paper is a companion to Vandenbergh et al. and Watts et al.
in Ecotoxicology and Environmental Safety, Volume 54, No. 2,
February 2003 on pp. 216–222 and 207–215, respectively, and to
Segner et al. in this issue.
*Corresponding author. Centre for Fish and Wildlife Health,
Institute of Animal Pathology, University of Bern, L.anggass-Str.
122, CH-3012 Bern, Switzerland. Fax: +41-31-631-2611.
E-mail address: [email protected] (H. Segner).
i.e., natural and man-made substances that interfere
with the endocrine system of vertebrates (Colborn and
Clement, 1992; Sharpe and Skakkebaek, 1993; Kavlock
et al., 1996; Ankley et al., 1998; Tyler et al., 1998; Vos
et al., 2000). EDCs have the potential to modulate or
disrupt the synthesis, secretion, transport, binding,
action, or elimination of endogenous hormones in the
body and consequently to affect homeostasis, development, reproduction, and behavior of organisms. Evidence from field studies, particularly of the aquatic
environment, suggests a relationship between exposure
to environmental hormones or hormone mimics and the
0147-6513/03/$ - see front matter r 2002 Elsevier Science (USA). All rights reserved.
PII: S 0 1 4 7 - 6 5 1 3 ( 0 2 ) 0 0 0 3 9 - 8
H. Segner et al. / Ecotoxicology and Environmental Safety 54 (2003) 302–314
manifestation of developmental and reproductive alterations in exposed organisms (e.g., Bryan et al., 1986;
Purdom et al., 1994; Matthiessen et al., 1995; Giullette
et al., 1996; Janssen et al., 1997; Lye et al., 1997; Jobling
.
et al., 1998; Oberdorster
and Cheek, 2000).
Hormones have a critical function in the regulation of
internal homeostasis in an organism and for its
acclimation to environmental factors. Since different
species inhabit specific ecological niches and environments, it may be expected that, although hormone and
receptor molecules are conserved in the animal kingdom, functional species specificities do occur. In view of
this, the possible extrapolation of endocrine-disrupting
effects from one species to another or to other
taxonomic groups is a crucial issue for the development
of test strategies for EDCs. In particular, it is not known
if compounds with endocrine activity in vertebrates are
also active in invertebrates. Most studies investigating
endocrine disruption in the aquatic environment have
focused on teleost fish, whereas invertebrates have
received little attention, despite the fact that they
constitute 95% of all living species and play an essential
role in the functioning and health of aquatic ecosystems.
Most notable among environmental EDCs are substances with steroid hormone activity, particularly with
estrogenic activity. Steroid substances are a group of
phylogenetically conserved hormones that act through a
common mechanism; i.e., they bind to cytoplasmic and
nuclear receptors and the ligand–receptor complex
subsequently activates transcription of steroid-responsive genes. In vertebrates, sex steroids have an ‘‘organizational’’ role in sexual differentiation during
development and an ‘‘activational’’ role in the sexual
cycle of mature, adult animals (Giulette et al., 1995;
Bigsby et al., 1999). The endocrine system of invertebrates relies largely on peptides and ecdysteroids.
Estrogens are also present; however, their physiological
role is not yet fully understood (deFur et al., 1999).
Assessment of the potential effects of environmental
estrogens, both natural estrogens and xenoestrogens, on
animal development and reproduction is complicated
due to (1) life stage-specific differences in the sensitivity
to steroid action, (2) the possible translation of subtle
physiological alterations during development into irreversible reproductive dysfunction of the adult organism,
(3) the pleiotropic actions of sex steroid hormones
within the organism, and (4) species and taxa specificities of steroid action. As mentioned above, the
assessment of endocrine-disrupting effects of environmental estrogens is difficult particularly for invertebrates because of the insufficient knowledge on the
physiological role of estrogens in these animal groups
.
(deFur et al., 1999, Oberdorster
and Cheek, 2000).
Current regulatory test guidelines for the assessment
of ecotoxicity are directed at measuring integrative
endpoints such as survival, growth, or reproduction.
303
This approach, however, makes it difficult to decide
whether an organism effect, e.g., an alteration of
reproduction, is due to endocrine activity of the test
compound. The challenge to researchers is to develop test
protocols that can answer the question whether a
developmental or reproductive effect induced by test
compound is in fact caused by the endocrine activity of
the substance (Tattersfield et al., 1997; CSTEE, 1999;
Fenner-Crisp et al., 2000; Huet, 2000). For the purpose of
EDC testing, the modification of existing guidelines of the
development of new guidelines is preferable since it would
support the acceptance and applicability of the tests.
The current article presents a summary of the findings
from the EU-funded research project ‘‘Identification of
Endocrine Disrupting Effects in Aquatic Organisms
(IDEA).’’ The objectives of the project were to examine
(a) what parameters and endpoints allow the detection
of endocrine-mediated developmental and reproductive
effects of (xeno)estrogens in life cycle- and life stagespecific toxicity tests with the zebrafish Danio rerio, a
small laboratory fish used in many ecotoxicity test
guidelines, and (b) whether substances that act as
estrogens in vertebrates may also adversely affect the
development, differentiation, and reproduction of aquatic invertebrates. The invertebrates examined in the
IDEA project included the coelenterate Hydra vulgaris,
the insect Chironomus riparius, the crustaceans Hyalella
azteca and Gammarus pulex, and the mollusc Lymnaea
stagnalis. As estrogenic test compounds, the synthetic
estrogen 17a-ethynylestradiol (EE2) and the xenoestrogens bisphenol A (BPA) and octylphenol (OP) were
selected. In addition, the biocide tributyltin (TBT) was
tested on molluscs, since this chemical has been linked
with endocrine disruption (imposex) in this animal group.
2. Assessment of developmental and reproductive
alterations related to endocrine disruption in long-term
biotests using the zebrafish
Much of the evidence for endocrine disruption in
wildlife populations has been derived from studies on
aquatic organisms, and, therefore, fish have been
recommended for the development of tests for EDCs
(Arcand-Hoy and Benson, 1998; Tyler et al., 1998;
OECD, 1999; Fenner-Crisp et al., 2000; Huet, 2000;
Harries et al., 2000). Full life cycle tests provide an
integrative approach to assess ecologically relevant
effects of chemicals on fish growth and reproduction,
but in order to recognize endocrine activity of the test
compound as the cause of developmental and reproductive alterations, it might be necessary to develop
modifications or enhancements of the existing test
protocols. In addition, since full life cycle tests are
laborious and costly, the question arises whether
‘‘sensitive windows’’ or ‘‘critical periods’’ do exist in
H. Segner et al. / Ecotoxicology and Environmental Safety 54 (2003) 302–314
304
the life cycle of zebrafish that would allow the detection
of the developmental and reproductive effects of EDCs
by means of short-term, life stage-specific exposures
instead of full life cycle experiments (see Table 1).
In the current project, the aim was to identify
developmental and reproductive disturbances related
to chronic exposure of zebrafish to compounds with
estrogenic activity. The zebrafish was chosen because
there are already test protocols in place, including
OECD guidelines (OECD, 1992, 1993, 1998) that
recommend zebrafish for chemical toxicity assessments.
The zebrafish has a relatively short life cycle of about 4
months and, in the laboratory, can be stimulated to
breed throughout the year. For this reason, zebrafish are
suitable for assessing toxic effects of chemicals on
development and reproduction (Nagel, 1993; Nagel
and Isberner, 1998; Andersen et al., 2000). As a positive
reference test compound, the synthetic estrogen EE2 was
selected. The effects induced by EE2 were compared
with the response pattern induced by exposure to the
xenoestrogens BPA and OP. The effects of estrogen
exposure on zebrafish were assessed by measuring
growth, survival, fecundity (number of eggs spawned
per female), fertilization rate, and time of first spawn.
These organism-level endpoints were combined with the
measurement of the biochemical indicator vitellogenin,
and with the analysis of gross morphology and
histopathology of the gonads. The synthesis of vitellogenin is under control of estrogens, and this protein is
acknowledged as a good biomarker of estrogenic
exposure in fish (Sumpter and Jobling, 1995). Quantification of vitellogenin in body homogenates of juvenile
Table 1
Summary of Effect Concentrations of Zebrafish, Danio rerio, Exposed to EE2 and BPA
Species
Response criterion
Zebrafish,
Danio rerio
Vitellogenin
Gonad histology
Survival
Juvenilea growth
Exposure period
From fertilization to
adult stage
Time to spawn
Mating behavior
Egg number per female
Fertilization succes
Hatching rate of offspring
Zebrafish,
Danio rerio
Significant induction
X1.67 ng/L
Changes X3 ng/L
No effect
Significant reduction
X1.67 ng/L
Significant delay X1.67 ng/L
Altered at X1.67 ng/L
Significant reduction
X1.67 ng/L
Significant reduction
X1.67 ng/L
No effect
Significant induction
X375 mg/L
Changes X375 mg/L
No effect
Significant reduction
X1500 mg/L
Significant delay X1500 mg/L
Altered at 1500 mg/L
Significant reduction at
1500 mg/L
Significant reduction at
1500 mg/L
No effect
0–3 dpf
No effect
Not determined
Survival
Juvenilea growth
Time to spawn
Mating behavior
Egg number per female
Fertilization success
Zebrafish,
Danio rerio
Effect Conc. BPA
Survival
Juvenilea growth
Time to spawn
Mating behavior
Egg number per female
Fertilization success
Zebrafish,
Danio rerio
Effect conc EE2
Not determined
0–21 dpf
Vitellogenin
Survival
Juvenilea growth
Time to spawn
Mating behavior
Egg number per female
Fertilization success
No effect
Significant induction X3 ng/L
0–75 dpf
42–75 dpf
No effect
No effect
Significant delay X3 ng/L
Altered at 3 ng/L and higher
Significant reduction X3 ng/L
Significant reduction X3 ng/L
Not determined
Note. The indicated test compound concentrations correspond to nominal concentrations. dpf, days postfertilization.
a
In the current experiments, the juvenile period went to 75 dpf (final bisexual gonad differentiation).
H. Segner et al. / Ecotoxicology and Environmental Safety 54 (2003) 302–314
zebrafish or in plasma of adult fish was achieved by a
homologous, competitive ELISA (Fenske et al., 2001).
Histology was included as an additional endpoint, since
a number of reports have demonstrated that gonad
histology of fish responds sensitively to exposure to
EDCs, with structural alterations ranging from the
appearance of ovo-testes to malformations of gametes
and stromal tissue or the increased frequency of atretic
oocytes. Histopathologic examinations may provide
insight into the nature of reproductive impairments
(Gimeno et al., 1996; Jobling et al., 1996, MilesRichardson et al., 1999, Metcalfe et al., 2000). At the
fourth meeting of the OECD Task Force on Endocrine
Disrupter Testing and Assessment (EDTA) in Paris on
May 12, 2001, it was generally agreed that vitellogenin,
in combination with gonad gross morphology and
histopathology, should be adopted as core endpoints
in the assessement of estrogen-active compounds.
2.1. Development of zebrafish
A prerequisite to correct interpretation of findings from
experiments on EDC exposure of zebrafish is knowledge
of the normal developmental and reproductive patterns of
this species. The zebrafish spawns in the morning when
the light switches on (Goolish et al., 1998). Each female
spawns up to 60 eggs every 2–3 days, depending on the
age of the individual (Eaton and Farley, 1974). Hatching
of the eggs occurs approximately 3 days after fertilization.
During development, zebrafish pass through a protogynic
stage, when all fish, irrespective of genetic sex, display
primary oocytes in the gonads (Takahashi, 1977; Maack
and Segner, 2002). This stage was reached in the present
experiments 40–45 days postfertilization (dpf); however, it
must be emphasized that the absolute time in days
required by zebrafish to reach a specific life stage may
vary widely among laboratories, depending on the culture
conditions and the zebrafish strain used.
After the protogynic stage, between 40% and 60% of
the juvenile fish undergo a transition of the ovary-like
gonad to a testis-like gonad, whereas the remaining
individuals continue to show an ovary-like morphology
of the gonad. In the authors’ laboratory, the fish
completed the process of morphological gonad differentiation between 65 and 75 dpf. From this stage on,
ovaries and testes display progressive growth and
maturation until onset of spawning. The female:male
ratio of zebrafish ranges around 50:50; however, it can
vary greatly among experimental cohorts. The spontaneous occurrence of intersex gonads containing both
oocytes and sperm cells is less than 1% in adult zebrafish
(Andersen et al., 2000, Maack and Segner, 2002).
2.1.1. Experiments with zebrafish: full life cycle test
Zebrafish were exposed to the model (xeno)estrogens
via water during a complete life cycle, from the fertilized
305
egg up to the sexually mature stage. The concentrations
of the test compounds were selected in order to allow
the observation of a concentration–response relationship: 0.05–10 ng EE2 L1, 94–1500 mg BPA L1, and
1.2–38 mg OP L1 (the values given are nominal concentrations; the analytically determined actual concentrations were close to the nominal ones). The test
concentrations of EE2 are environmentally realistic,
whereas for BPA and OP, the test concentrations are
higher than those typically found in European surface
waters.
In zebrafish exposed to EE2, mortality was not
enhanced compared to controls, but growth of juveniles
(up to day 75 pf) was reduced at the higher concentrations. Life cycle estrogen exposure resulted in a delayed
developmental onset of spawning, in a reduction of the
number of eggs produced per female, and in a reduction
of the fertilization success. Whereas the decline in
fertilization success was clearly expressed in all experiments, the response of the two other parameters—egg
number of female, time of first spawning—could be
obscured by high variation among individual or experimental groups. Plasma vitellogenin levels in both male
and female zebrafish became significantly elevated with
estrogen treatment. In the case of EE2, the lowestobserved-effect concentration (LOEC) value for changes
of the fertilization success and for VTG induction were
both at 1.67 ng EE2 L1 (nominal concentration),
whereas in the case of BPA, the LOEC for vitellogenin
was lower (375 mg BPA L1) than the LOEC for the
fertilization capability (1500 mg BPA L1). These findings indicate that in chronic exposure experiments with
zebrafish, the sensitivity of reproductive parameters
measured at the organism level is comparable to that of
molecular markers such as vitellogenin, although some
variation dependent on the compound tested may occur.
Thus, the primary function of molecular endpoints in
the chronic life cycle exposure test with zebrafish may be
to increase the specificity rather than the sensitivity of
the test.
For some fish species, the occurrence of intersex or a
shift of the gonadal sex ratio (i.e., the ratio of ovaries
and testes) has been described after prolonged estrogen
exposure (Gray and Metcalfe, 1997; Nimrod and
Benson, 1998; Scholz and Gutzeit, 2000). It was therefore decided to explore whether the gonadal morphology of zebrafish is sensitive to estrogen exposure. A
treatment-related increase in the number of ovotestes
was not observed in the current experiments, but chronic
EE2 exposure led to an increased percentage of ovariantype gonads. Congruent findings have been reported by
. et al. (2000). However, recovery experiments, after
Orn
termination of the estrogen treatment, reveal that the
skewed ovary/testis ratio of the previously exposed
population returns to a ratio comparable to that of the
controls. This observation may indicate that there
306
H. Segner et al. / Ecotoxicology and Environmental Safety 54 (2003) 302–314
occurs a transient modification of the gonadal differentiation process rather than a permanent and irreversible effect.
Exposure of zebrafish to 10 ng EE2 L1 completely
inhibited reproduction; neither mating behavior nor
spawning took place. The inhibition of zebrafish
reproduction by 10 ng EE2 L1 was virtually irreversible: when fish reared from the fertilized egg until the
adult stage under exposure to 10 ng EE2 L1 were
transferred into noncontaminated water, spawning was
resumed after 3 months, but other reproductive parameters did not or only partly recovered; vitellogenin
levels remained significantly elevated, gonad morphology continued to show structural aberrations, and, most
importantly, fertilization success remained below 5%. It
is not currently known whether this low fertilization
success is due to impaired quality of the sperm or
the eggs.
Treatment of zebrafish with BPA and OP led to
identical alterations as described for EE2, with the
only difference that the effect threshold concentrations
were in the microgram per liter range instead of
nanograms per liter. It appears that the response pattern
observed after the one-generation exposure of zebrafish
to EE2, OP, and BPA is indicative of an estrogenic
mode of action of a test compound. This ‘‘estrogenic
pattern’’ differs from developmental and reproductive
changes as induced by nonestrogenic xenobiotics. In
zebrafish life cycle experiments with compounds such as
4-chloroaniline (Bresch et al., 1990), 2,4-dichloroaniline
(Ensenbach and Nagel, 1997), 2,3,7,8-tetrachlorodibenzo-p-dioxin (Wannemacher et al., 1992), or 1,2,3trichlorobenzene (Roex et al., 2001), the reproductive
parameter most affected was the number of eggs
produced per female, while fertilization success or
the time of first spawning were not altered. Thus, a
reduced fertilization success, particularly in combination
with elevated vitellogenin levels and modifications of
gonadal differentiation, appears to be indicative of
an estrogenic mode of action of the test compound
on zebrafish.
2.2. Experiments with zebrafish: partial life cycle test
Life cycle assays are very challenging to perform, and
there is a practical need to develop more pragmatic
partial life cycle tests or in vivo screens. Examples of
those assays include the gonadal recrudescence test
(Huet, 2000), or the 6-week reproductive performance
test developed for fathead minnow (Harries et al., 2000).
The authors’ approach to this problem was to search for
a period in zebrafish development that is particularly
sensitive to estrogens. As stated by Bigsby et al. (1999),
development is epigenetic, with endocrine signals
coordinating differentiation during precise times in
development and at specific dose ranges. This means
that irreversible effects on sexual differentiation and
function might be possible by exposure to environmental estrogens during a critical period in development. Thus, exposure of zebrafish during such a critical
period may predict the outcome of estrogen exposure
during a full life cycle experiment.
In order to determine if a critical period of estrogen
sensitivity exists in zebrafish ontogenesis, zebrafish were
exposed during different developmental stages: (a) from
fertilization until the end of the embryonal period (start
of hatching at 3 dpf), (b) from fertilization until 21 dpf
(phase of undifferentiated gonads), (c) from fertilization
until 42 dpf (phase of protogynic ovaries), (d) from
fertilization until 66–72 dpf (completion of male/female
differentiation), and (e) from fertilization until the
reproductive age, i.e., during the full life cycle. In
addition, experiments were performed where fish were
exposed not from fertilization onward, but only during
the period of bisexual morphological differentiation of
the gonads (in the authors’ system, this occurs between
42 and 75 dpf). At the end of the exposures, the animals
were transferred to noncontaminated water and were
reared under control conditions until the adult stage.
Then, the reproductive parameters (start of spawning
and mating behavior, number of eggs per female,
fertilization success, hatching success of embryos,
embryo survival) were recorded for a 3-week period.
Additionally, gonad histology and vitellogenin levels
were studied. The test compounds and concentrations
used in the partial life cycle tests were the same as those
applied in the full life cycle experiments.
Estrogenic exposure of zebrafish during the embryonal (0–3 dpf), larval (0–21 dpf), or early life stage
(0–42 dpf) period remained without lasting effect on
reproductive performance, vitellogenin levels, and gonad differentiation at the adult stage. These results
indicate that the suggestion of CSTEE (1999) to use fish
early life stage tests as an estrogenic screen is not
suitable with zebrafish.
Exposure of zebrafish during the period of sexual
differentiation (from fertilization until 75 dpf or only
from Days 42 to 75 pf) resulted in changes of the
reproductive performance of the adult fish. The induced
effects were identical to those observed in the full life
cycle exposures of zebrafish (reduced fertilization capability, delayed start of spwaning, elevated vitellogenin
levels, altered gonadal differentiation), but the effect
strength was weaker than in the full life cycle experiments. For instance, whereas in the life cycle study a
significant reduction of fertilization capabilities occurred
at 1.67 ng EE2 L1, it took 3 ng EE2 L1 in the
42–75 dpf exposure to obtain a significant change of
fertilization capabilities. The results are promising that
partial life cycle exposure of zebrafish during the period
of final gonad differentiation may substitute for the full
life cycle test for estrogenic substances.
H. Segner et al. / Ecotoxicology and Environmental Safety 54 (2003) 302–314
3. Assessment of (xeno)estrogen-induced developmental
and reproductive alterations in invertebrates
For the invertebrates (Tables 2 and 3), a major
problem in establishing suitable endpoints to assess
endocrine disruption is the insufficient knowledge
regarding the hormone systems of these animal taxa.
The complexity of reproductive systems and life
histories of invertebrates, including metamorphosis,
diapause, and regeneration, has resulted in the evolution
of endocrine control systems significantly different from
those of the vertebrates (DeFur et al., 1999; Pinder et al.,
1999; LaFont, 2000) and it is not clear therefore if
307
vertebrate endocrine disrupters, particularly estrogens
and estrogen mimics, also present a risk to invertebrates.
Although compared with the vertebrates knowledge of
invertebrate endocrine systems is limited, for three
groups (crustaceans, insects, and molluscs), the systems
are relatively well understood. Endocrine control
systems in the development, growth, and reproduction
of invertebrates often involve peptidic hormones. For
instance, ovulation and egg-laying behavior in L.
stagnalis are regulated by a neurosecretory peptide, the
egg-laying hormone (Geraerts et al., 1991). In crustaceans and insects, invertebrate-specific steroids (the
ecdysteroids) and terpenoids (the juvenile hormones)
Table 2
Summary of effect concentrations for Hydra vulgaris, Gammarus pulex, and Chironomus riparius exposed to EE2 and BPA
Species/life stage
Response criterion
Exposure
period
Effect conc EE
Effect conc BPA
H vulgaris
(Zurich strain,
male clone)
Mortality/LC50
Regeneration of DR/
degeneration of polyp
96 h
72 h
H. vulgaris (US
strain, separate
sexes)
No. of testes/oocytes
6 weeks
3.78 mg/L
Regeneration inhibited at 320 mg/
L; no effect on degeneration up
to 1.6 mg/L; No effects at low
(10 ng–58 mg/L) conc
Significant ðPo0:05Þ reduction
at 0.5 mg/L
6.91 mg/L
Regeneration inhibited at 1.0 mg/
L. No effect on degeneration up
to 4.6 mg/L; No effects at low
(21 ng–42 mg/L) conc
—
Sperm activity
6 weeks
Significant ðPo0:05Þ reduction
at 0.5 mg/L
—
Mortality/LC50
Direct precop separation
Indirect precop separation
240 h
24 h
24 h
0.84 mg/L
No effect (10 ng/L–3.7 mg/L)
No effect (10 ng–3.7 mg/L)
Reforming of pairs
24 h
Population structure/
recruitment
100 days
Significant effects at high
(1.0–3.7 mg/L)
Significant effects ðPo0:05Þ on
population size; increased
recruitment at 1–10 mg/L; mean
control popn=169; 1 mg/L=385;
10 mg/L=411
42:1 female sex ratio; 2nd sex
characters not affected
1.49 mg/L
No effect (10 ng/L–8.4 mg/L)
Significant effect ðPo0:05Þ at
8.4 mg/L
Only affected at high
concentration (0.83 mg/L)
Not tested
Mortality/LC50
Emergence/reproduction
240 h
2 life cycles
Eggs
Moulting/development
ca. 20 days
4th instar (from
above test)
Mouthpart deformities
—
G . pulex juvenile
Adult pairs
Neo., juv., adults
C. riparius
2nd instar
1st instar
8.83 mg/L
1st generation: X90% adults
emerged in all treatments; little
effect on EmT50; effects on egg
prodn not dose related;
2nd generation: significantly
ðPo0:05Þ more adults emerged
at 50 ng/L. Adult sex ratio
differed from 1:1 at 10 and
50 ng/L;
Moulting delayed/wet weight
reduced at 1 mg/L
Significant deformities of
mouthparts noted at 10–10 mg/L.
Little or no effect at high conc
Note. The indicated test compound concentrations correspond to nominal concentrations.
11.51 mg/L
1st generation: X90% adults
emerged in all treatments. Little
effect on EmT50. Effects on egg
prodn not dose related;
2nd generation: X85% of adults
emerged—no emergence at
10.4 mg/L; significant effects
ðPo0:05Þ on EmT50—delayed at
78 ng–0.75 mg/L
Moulting delayed/wet weight
reduced at 1 mg/L
Less incidence of deformities
than with EE Effective conc.
similar (10 ng–10 mg/L)
308
H. Segner et al. / Ecotoxicology and Environmental Safety 54 (2003) 302–314
Table 3
Summary of effect concentrations for Hyalella azteca and Lymnaea stagnalis exposed to EE2 and TBT
Species/Life stage
H. azteca
P generation
F1 generation
Gametogenesis
until adult
Post F2-generation
Gametogenesis
until adult
F3 generation
Gametogenesis
until adult
L. stagnalis
Egg masses
Juvenile (40 days)
Response
criterion
Exposure
period
Effect conc EE2
Effect conc TBT
Secondary sex
characteristics
Secondary sex
characteristics
4 weeks
Male and female: no effect on length of
antennae and gnathopods (0.1–10 mg/L)
Male: smaller 2nd gnathopods
(0.1–0.32 mg/L)
Male and female: no effects on antennae
(0.1–10 mg/L)
Not tested
6 weeks
Histopathology
15 weeks
Male: disturbed gonadal development
(0.1–10 mg/L)
Not tested
Sex ratio
5 weeks
No effect on sex ratio (0.1–10 mg/L)
Not tested
Hatching/
development/
protein
metabolism
3 weeks
Delayed hatching (1000 ng/L)
Disturbed hatching (320 ng/L)
Disturbed hatching (500–1000 ng/L)
Deformations developing snails
(32–320 ng/L)
Mortality/
histopathology/
Ca metabolism
3 weeks
Deformations developing snails
(100–1000 ng/L)
Altered protein pattern (50–500 ng/L)
Disturbed development secretory cells
digestive gland (500–1000 ng/L)
Central cavity prostate gland
(100 ng/L)
Decreased calcification of shell
(1000 ng EE/L)
Adult (80 days)
Adult (80 days)
Juvenile (40 days)
Egg masses
Juvenile (40 days)
2nd generation egg
masses
Not tested
Mortality/
histopathology/
Ca metabolism
3 weeks
Decreased calcification of shell
(1000 ng/L)
Behavior/
physiology
Several
hours
No effects
Protein
metabolism/
growth
10 weeks
Reduced growth hatchlings
(50–500 ng/L)
Egg laying/
hatching/protein
metabolism
10 weeks
7-day LC50: 4.72 mg/L
21-day LC50: 0.71 mg/L
Vacuolization and enlargement
prostate gland (0.23 mg/L)
Disturbed development secretory
cells digestive gland (0.32 mg/L)
Decreased calcification of shell
(0.23–3.2 mg/L)
7 day LC50: 5.94 mg/L
21 day LC50: 0.71 mg/L
Vacuolisation prostate gland
(0.32 mg/L)
Necrosis prostate gland, digestive
gland (3.2 mg/L, 2 weeks)
Effects on spermduct, pars
contorta oviductus (3.2 mg/L, 2
weeks)
Decreased calcification of shell
(0.23–3.2 mg/L)
Eversion penial complex (11 mg/L)
Not tested
Altered protein pattern (50- 500 ng/L)
Altered protein pattern (50–500 ng/L)
Not tested
Increased egg laying (more egg masses,
more eggs per egg mass)
Delayed hatching (500 ng/L)
Not tested
Detachment from substrate
(50–100 ng/L)
Note: The indicated test compound concentrations indicated correspond to nominal concentrations.
H. Segner et al. / Ecotoxicology and Environmental Safety 54 (2003) 302–314
regulate molting, differentiation, and metamorphosis,
but they are also involved in reproductive processes such
as ovulation, spermiogenesis, and vitellogenesis. The
structural similarity of vertebrate estrogens and ecdyson
points to the possibility that estrogenic compounds
could interfere with endogenous steroids in invertebrates. Zou and Fingerman (1997) have observed that
certain polychlorinated biphenyls (PCBs) that have
estrogenic effects on vertebrates, acted as moulting
inhibitors in the water flea Daphnia magna. The authors
suggest that the tested PCBs act as antagonists of
endogenous ecdysteroids by forming inactive complexes
with ecdysteroid receptors. In addition to ecdysteroids,
the vertebrate steroids estrogens and testosterone have
also been found in invertebrates, but it is not clear
whether they have a functional role and through which
receptors they become active (Fingerman et al., 1993;
Hood et al., 2000; LaFont, 2000). While the ability of
many invertebrates to synthesize vertebrate-type steroids is questionable, the gonads of molluscs are clearly
able to synthesize estrogen and testosterone de novo. It
has been suggested that the TBT-induced imposex in
certain mollusc species is caused by a disturbance of
estrogen biosynthesis (Spooner et al., 1991).
In the current project, representative species of
molluscs, crustaceans, and insects were examined. In
addition, the cnidarian H. vulgaris was included as test
organism. Cnidarians are primitive invertebrates representing an evolutionary line present before the divergence of the protostomes including most invertebrate
groups, and deuterostomes including the vertebrates.
Their endocrinology may therefore represent features
that are conserved by both groups.
Assessing the potential endocrine-disrupting effects of
chemicals for which the natural receptors may not even
be present in the investigated target species requires a
more general approach relying on integrative organismic
response criteria such as behavior, development, reproduction, and full life cycle characteristics, instead of
using specific molecular endpoints of endocrine disruption. Consequently, the first aim of the current project
was to determine if estrogens have any effects on
developmental and reproductive parameters in aquatic
invertebrates, without initially considering whether an
observed effect in fact results from the disturbance of
the endocrine system. The parameters examined in the
project included the regeneration of dissected H. vulgaris
preparations into complete polyps, the sexual development of H. vulgaris polyps (Pascoe et al., 2002), the
separation and formation of precopulatory pairs in G.
pulex, and the effect of EE2 on the population structure
of this species. Several tests were performed with the
insect C. riparius, to assess chemical effects on moulting
and mouthpart structure in the larvae and the emergence and reproduction of adults. For the crustacean H.
azteca, full life cycle and multigeneration exposures were
309
performed to explore effects on various developmental
and reproductive parameters. Finally, the mollusc L.
stagnalis was investigated; here, as with H. azteca, a
number of developmental and reproductive parameters
(growth, egg-laying, hatching) were investigated. In
addition, suborganismic parameters related to hormonal
processes such as calcium metabolism, gonad histopathology, and vitellogenin synthesis were examined in
order to establish a more direct, mechanistic link to
estrogenic action.
In the acute precopula separation tests with G. pulex
(Watts et al., 2001a) no indication of effects at low
exposure concentrations of (xeno)estrogens were recorded (Table 2). Significant effects on G. pulex at lower
concentrations were identified in the longer-term population study (Watts et al., 2002a). Significant increases in
recruitment and population size were recorded at 1 and
10 mg L1 EE2. Mean population size increased in all
treatment groups at the end of the 100-day exposure,
primarily attributable to recruitment with X70% of the
animals identified as neonates and juveniles (1.5–6.0-mm
length). However, at 1 and 10 mg EE2 L1, the respective
mean population sizes of 385 and 411 animals were
significantly ðP ¼ 0:018Þ greater than the control mean
of 169 animals. In addition, the sex ratio was biased by
42:1 in favor of females in the treated groups (100 ng
EE2 L1–10 mg EE2 L1). Additional parameters such as
the number of precopula pairs/ovigerous females and
measurement of secondary sexual characteristics showed
no significant differences ðP40:05Þ between treatments.
The estrogen treatment-related increases in population
size may be partially explained by an increased rate of
female sexual maturation in the exposed animals, a view
supported by the results from several other crustacean
studies (Sarojini et al., 1986, 1990).
In the longer-term tests with C. riparius some effects
on emergence times and adult numbers were associated
with the second generation of exposed animals at
environmentally relevant concentrations (Watts et al.,
2001b). However, little effect was seen in the first
generation. In addition, there was some indication that
the test chemicals EE2 and BPA induced different
responses in the test animals. Exposure to EE2 resulted
in significant effects on adult numbers and sex ratio;
however, in the case of BPA, neither of these parameters
was affected but emergence times were delayed. Results
from these experiment do suggest that there has been
some disruption of the normal development process
but inconsistencies in the data do not support firm
conclusions.
The assessment of mouthpart structure in C. riparius
provided the clearest indication of an effect at low
concentrations in this species. Moulting was affected
only at high, environmentally unrealistic concentrations
(1 mg L1), but mouthpart deformities were mainly seen
in the range 10 ng L1–10 mg L1 of EE2 and BPA. In
310
H. Segner et al. / Ecotoxicology and Environmental Safety 54 (2003) 302–314
contrast to the emergence test, induction of mouthpart
deformities followed a similar pattern for both chemicals, with deformity primarily associated with the
mentum. In accordance with the greater estrogenicity
of EE2, the incidence of deformity was higher with this
chemical than with BPA (Watts et al., 2002b). Other
groups (Meregalli et al., 2001) have published data
indicating that the xenoestrogen 4-nonylphenol induced
mentum deformities in this species, but Meregalli and
Ollevier (2001) did not find deformities with EE2
(1–100 mg L1). The results seem to suggest that there
may be some interaction betweeen estrogenic chemicals
and the ecdysteroid receptors of C. riparius. In view of
this, mouthpart deformity may be worth further investigation as a possible biomarker of estrogenic exposure.
With H. azteca, full life cycle (4–6 weeks, from
gametogenesis or hatch, respectively, until adulthood)
and multigeneration exposure (15 weeks) led to several
organizational effects, mainly smaller male second
gnathopods and disturbed gonadal development (Vandenbergh et al., 2002) but similar to what has been
observed with C. riparius, only in the second generation.
Promising results were also obtained for L. stagnalis.
Exposure to EE2 starting at the juvenile stage and on
through sexual maturity caused an increase in egg-laying
(more egg masses with more eggs) but hatching of these
egg masses was disturbed. Life cycle exposure to EE2
starting with fresh egg masses showed a decrease in
growth in hatched snails and an alteration in the protein
metabolism, which may be linked with a vitellogeninlike protein. Also, a different protein metabolism was
seen in egg masses after exposure to EE2, and exposure
of adult organisms to EE2 evoked changes in protein
metabolism, including the induction of a vitellogeninlike protein. Similar observations have been reported by
Blaise et al. (1999) and Gagne! et al. (2001). At the
present stage of knowledge, however, it is too premature
to suggest that vitellogenin is an estrogen biomarker for
molluscs, since the hormonal regulation of this protein is
only partly known and the dose–response relationship
of estrogenic induction remains to be established.
Even in the absence of established biomarkers, shortterm experiments with L. stagnalis may be already
possible for some endpoints. A dose–response effect was
found on hatching of egg masses and development of
embryos after exposure to EE2 and also TBT, a
compound used for comparison since it is a known
EDC for molluscs. Short-term exposure of juveniles (40day-old) and adults (80-day-old) to both EE2 and TBT
induced a decreased calcification of the shell and
histological aberrations of the reproductive tract,
although the expression of histological effects was
greater in juvenile organisms. Generally, the effects of
EE2 exposure were more pronounced when L. stagnalis
was exposed early in the life cycle (egg masses, new
hatchlings, juvenile stage). Histological alterations of
reproductive organs were mainly seen in the prostate
after exposure to TBT and may be linked to a
neuroendocrine activity of TBT compounds, whereas
no clear-cut effects occurred under EE2 exposure.
The invertebrate test results do indicate that the
synthetic estrogen EE2 and xenoestrogens such as BPA
bring about different effects in exposed organisms. For
example, in the C. riparius emergence test, significant
effects were associated with adult numbers and sex ratio
of animals exposed to EE2 while in the case of BPA,
emergence times were delayed but the other parameters
were not affected. However, the induction of mouthpart
deformities in C. riparius followed a similar pattern for
both chemicals, with deformity mainly associated with
the mentum. In agreement with the greater estrogenicity
of EE2, the incidence of deformity was higher with this
chemical than with BPA. These data are typical of the
results obtained with the invertebrates in that they
demonstrate an inconsistency in organism response to
the two test chemicals which, in vertebrates at least,
bring about effects via interaction with the same
target—the estrogen receptor.
4. Recommendations for test procedures with fish and
aquatic invertebrates
One goal of the project was to derive from the
experimental results recommendations on how to
enhance existing toxicity test procedures with respect
to the assessment of estrogen-active substances.
Of the invertebrates used in the IDEA program,
standardized regulatory tests are available for two
species—C. riparius and H. azteca. No standard tests
exist for H. vulgaris, G. pulex, or L. stagnalis, although
these species are often used in ecotoxicological research.
In the case of chironomids, standard tests used by
ASTM, USEPA, Environment Canada, and OECD
utilize growth or emergence over an exposure period of
10–28 days. It is relatively easy to modify these test
protocols to include additional endpoints which may
provide an indication of effects on the endocrine system.
For example, a standard emergence test can be extended
to incorporate the effects on reproduction (number of
eggs produced) and subsequent effects on a second
generation of test animals. Also, the 10-day growth
assay could be adapted so that in addition to weight
determination, the mouthparts of fourth-instar larvae
are examined for any deformities at the end of the
exposure period.
For H. azteca, which is predominantly used in North
America, standard tests ranging from p10 to 30 days
are recommended by the ASTM, US EPA and
Environment Canada. The endpoints examined depend
on the test duration and include survival and growth in
short-term assays while exposure up to 30 days allows
H. Segner et al. / Ecotoxicology and Environmental Safety 54 (2003) 302–314
the assessment of reproductive capacity in terms of the
number of young produced. These tests would require
little modification to examine the effects of EDCs on
reproduction/development, as demonstrated by the
findings of this project. Although there is no standard
test for G. pulex, response criteria similar to those
described for H. azteca were used in a population study
of 100 days, with significant effects on population
structure noted at 1 and 10 mg L1 of EE2.
In view of the limited knowledge of invertebrate
endocrinology, life cycle testing is preferred in tests with
invertebrates. Except for L. stagnalis, where endpoints
such as hatching, histology, and protein metabolism
may be used for screening, none of the short-term tests
with the other species selected in the project suggest that
effects can be detected in short-term assays. The
possibility to detect an effect which may have resulted
from disruption of the endocrine system is greatest in
tests using the full life cycle where various processes
(development, growth, moulting, reproduction) are
controlled by the endocrine system and therefore
provide potential targets for disruption. Recognition
of this is provided by the recommendation that full life
cycle tests be adopted as the ‘‘gold standard’’ for
assessment of EDCs in invertebrates (Ingersoll et al.,
1999).
Although modification of existing test protocols to
include response criteria which are potential targets for
EDCs may be relatively easy to implement with
invertebrates, the data obtained during the IDEA
project, while suggestive of effects on reproduction
and development, do not provide unequivocal proof of
endocrine disruption. However, in the absence of
specific markers of exposure, integrative response
criteria (growth, reproduction, etc.) currently provide
the only alternative for the assessment of EDCs in
invertebrates.
Concerning fish, the need for a definitive, tier II test
on endocrine disruption has been widely recognized
(EDSTAC, 1998; OECD, 1999; CSTEE, 1999), however, questions concerning the most suitable species and
test endpoints for the asssessment of endocrine-related
effects are still to be resolved. The results of the present
project provide evidence that in a life cycle test with
zebrafish estrogenic disruption can be recognized from a
characteristic response pattern including reduced fertilization success, delayed start of spawning, elevated
vitellogenin levels and altered gonad differentiation. It is
an important finding of this project that an estrogenic
activity of a test compound is already indicated
from alterations of organism parameters such as
fertilization success, which are routinely measured in
life cycle tests, although confirmation by estrogenspecific endpoints such as vitellogenin induction is
required (admittedly, even the combination of, e.g.,
reduced fertilization success and elevated VTG does not
311
fully provide proof an estrogen receptor-dependent
mechanism of the altered reproductive performance,
however, a regulatory test cannot provide the ultimate
mechanistic verification of an observed effect). Technically, the measurement of estrogen-indicative endpoints
requires only minor modifications of existing test
protocols for the zebrafish. The suggested estrogenic
endpoints can be easily and rapidly measured, for
instance, in the case of vitellogenin, standard ELISA
techniques may be used (Fenske et al., 2001). Another
important finding is that a partial life cycle test
encompassing the period of final gonad differentiation
of zebrafish has the potential to substitute for the full life
cyle test when it comes to the evaluation of estrogenic
activities.
5. Conclusions
In the current study, the focus was on developmental
and reproductive changes resulting from chronic exposure of fish or aquatic invertebrates to exogenous
(xeno)estrogens. The chemical concentrations used in
the experiments were selected to approach environmentally realistic concentrations and to be clearly below the
chronic lethality levels—for instance, in zebrafish the
28-day LC50 for EE2 is 100 ng L1, whereas the life cycle
LOEC for the EE2-induced reduction of the fertilization
success is 1.67 ng L1. The results clearly demonstrate
that compounds with estrogenic activity are able to
impair at low concentrations the development and
reproduction of zebrafish as well as various developmental and reproductive parameters of aquatic invertebrates. Whereas in the case of zebrafish, there is good
evidence that the observed reproductive alterations are
in fact due to the hormonal activity of the test
compounds, it is too early to draw such a conclusion
with respect to the invertebrates.
A number of organizations and agencies are currently
trying to establish both screening (tier I) and definitive
(tier II) tests for endocrine disruptors (EDSTAC, 1998;
OECD, 1999; Fenner-Crisp et al., 2000; Huet, 2000).
Tier I screening is designed to detect a substance’s
potential for causing disruption, and tier II testing is
designed to provide definitive proof of a substance’s
ability to interact adversely with the hormone system in
the intact organisms and to disrupt hormone-regulated
physiological functions. In ecotoxicology, emphasis in
tier II test development is on protocols using small
laboratory fish, mainly medaka (Oryzias latipes), fathead minnow (Pimephales promelas), and zebrafish
(Danio rerio). Our data reveal how to assess the
disrupting effects of estrogenic compounds in a life
cycle test with zebrafish by a combination of molecular,
cellular, and organism endpoints. With three (xeno)estrogens (EE2, OP, BPA) tested to date in the chronic
312
H. Segner et al. / Ecotoxicology and Environmental Safety 54 (2003) 302–314
zebrafish test, the available database is small; however,
the fact that all three compounds induced an identical
response pattern suggests that this pattern in fact is
estrogen-specific for zebrafish.
With respect to the invertebrates, the results of the
project provide evidence that certain effect parameters
are responsive to exogenous estrogens; however, it is not
possible to conclude categorically that the effects are
hormone-mediated and result from an interaction with
the endogenous endocrine system of the invertebrates.
In addition, the effects often have no clear concentration–response relationship, and occur at rather high
concentrations (factor 1000 higher than in fish), which
suggests general toxicity rather than endocrine effects.
However, certain response criteria in invertebrates, e.g.,
mouthpart deformities and population structure were
affected at low concentrations, although currently the
mechanisms underlying the effects are not understood,
and therefore, the use of these endpoints to indicate an
estrogenic activity of a test substance is questionable.
Data from the G. pulex population study indicate that
significant effects can occur in relation to reproduction,
however, they cannot be predicted from the results of
short-term tests. Similarly, deformities of the mouthparts do not relate to effects on development or growth.
The results currently do not allow the establishment of
short-term assays or biomarker-type responses which
would indicate that the compound exerts an estrogentype activity in the animal. However, promising findings
with respect to the possibility of short-term in vivo
screens were obtained in the experiments with L.
stagnalis
As far as the invertebrates are concerned, it appears to
be not possible to identify critical periods of estrogen
sensitivity because the level of understanding of
invertebrate endocrinology is still inadequate. The
extent to which the endpoints measured in the experiments reflect an estrogenic mode of action remains
unclear. Even for the suborgansmic parameters such as
induction of vitellogenesis, the involvement of estrogen
is not established for invertebrates. The most promising
results that were seen in the invertebrate experiments of
this project (C. riparius mouthpart deformities, G. pulex
population changes) were produced by full life cycle
exposures, which vary from 20 to 100 days, depending
on the species used. At present, full life cycle experiments appear to be the most appropriate exposure
regime to reveal sublethal effects of environmental
estrogens on invertebrates.
Acknowledgments
The current project was financially supported by
European Commission Contract ENV4-CT97-0509.
References
.
Andersen, L., Bengtsson, B.E., Bjork,
M., Gessbo, A., Holbech, H.,
Hylland, K., Norrgren, L., Pedersen, K.L., Lundgren, A., Petersen,
.
G.I., Steinholz, A., Orn,
S., 2000. Zebrafish for Testing Endocrine
Disrupting Chemicals. TemaNord 555, Copenhagen.
Ankley, G., Michaich, E., Stahl, R., Tillit, D., Colborn, T., McMaster,
S., Miller, R., Bantle, J., Cambell, P., Denslow, N., Dickerson, R.,
Folmar, L., Fry, M., Giesy, J., Gray, L.E., Guiney, P., Hutchinson,
T., Kennedy, S., Kramer, V., LeBlanc, G., Mayes, M., Nimrod, A.,
Patino, R., Peterson, R., Purdy, R., Ringer, R., Thomas, P.,
Touart, L., van der Kraak, G., Zacharewski, T., 1998. Overview of
a workshop on screening methods for detecting potential (anti)estrogenic/androgenic chemicals in wildlife. Environ. Toxicol.
Chem. 17, 68–87.
Arcand-Hoy, L.D., Benson, W.H., 1998. Fish reproduction: an
ecologically relevant indicator of endocrine disruption. Environ.
Toxicol. Chem. 17, 49–57.
Bigsby, R., Chapin, R.E., Daston, G.P., Davis, B.J., Gorski, J., Gray,
L.E., Howdeshell, K.L., Zoeller, R.T., vom Saal, F.S., 1999.
Evaluating the effects of endocrine disruptors on endocrine
function during development. Environ. Health Perspect. 107
(Suppl. 4), 613–618.
Blaise, C., Gagn!e, F., Pellerin, J., Hansen, P.D., 1999. Determination
of vitellogenin-like properties in Mya arenaria hemolymph
(Sagenuay Fjord., Canada): a potential biomarker for endocrine
disruption. Environ. Toxicol. 14, 455–465.
Bresch, H., Beck, H., Ehlermann, D., Schlaszus, H., Urbanek, M.,
1990. A long-term toxicity test comprising reproduction and
growth of zebrafish with 4-chloroaniline. Arch. Environ. Contam.
Toxicol. 19, 419–427.
Bryan, G.W., Gibbs, P.E., Hummerstone, L.G., Burt, G.R., 1986. The
decline of the gastropod Nucela lapillus around South-West
England: evidence for the effect of tributyltin from antifouling
paints. J. Mar. Biol. Assoc. UK 66, 611–640.
Colborn, T., Clement, C., 1992. Chemically Induced Alterations In
Sexual Development: The Wildlife/Human Connection. Princeton
Scientific, Princeton.
CSTEE (European Committee on Toxicity, Ecotoxicity and Environment), 1999. Opinion on Human and Wildlife Health Effects of
Endocrine Disrupting Chemicals, with Emphasis on Wildlife and
Ecotoxicology Test Methods, Brussels.
deFur, P.L., Crane, M., Ingersoll, C., Tattersfield, L., 1999. Endocrine
Disruption in Invertebrates: Endocrinology, Testing and Assessment. Society of Environmental Toxicology and Chemistry,
Pensacola, FL, USA.
Eaton, R., Farley, R., 1974. Spawning cycle and egg production in
zebrafish, Brachydanio rerio, reared in the laboratory. Copeia 1,
195–203.
EDSTAC (Endocrine Disruptor Screening and Testing Advisory
Committee), 1998. Final Report. Executive Summary, Vols. 1, 2.
Washington, DC.
Ensenbach, U., Nagel, R., 1997. Toxicity of binary chemical mitures:
effects on reproduction of zebrafish (Brachydanio Rerio). Arch.
Environ. Contam. Toxicol. 32, 204–210.
Fenner-Crisp, P.A., Maciorowski, A.F., Timm, G.E., 2000. The
endocrine disruptor screening program developed by the US
Environmental Protection Agency. Ecotoxicology 9, 85–91.
Fenske, M., van Aerle, R., Brack, S., Tyler, C.R., Segner, H., 2001.
Development and validation of a homologous zebrafish (Danio
rerio hamilton-buchanan) vitellogenin enzyme-linked immunosorbent assay (ELISA) and its application for studies on estrogenic
chemicals. Comp. Biochem. Physiol. 129c, 217–232.
Fingerman, M., Nababushaman, K., Sarojini, R., 1993. Vertebratetype hormones in crustaceans—localization, identification and
functional significance. Zool. Sci. 10, 13–29.
H. Segner et al. / Ecotoxicology and Environmental Safety 54 (2003) 302–314
Gagn!e, F., Blaise, C., Salazar, M., Hansen, P.D., 2001. Evaluation
of estrogenic effects of municipal effluents to the freshwater
mussel Elliptio complanata. Comp. Biochem. Physiol. C 128,
213–225.
Geraerts, W.P.M., Smit, A.B., Li, K.W., Vreugdenhil, E., van
Heerikhuizen, H., 1991. Neuropeptide gene families that control
reproductive behavior and growth in molluscs. In: Osborne, M.N.
(Ed.), Current Aspects of the Neurosciences. MacMillan, London,
pp. 255–285.
Gimeno, S., Gerritsen, A., Browmer, T., 1996. Feminization of male
carp. Nature 384, 221–222.
Giulette, L.J., Crain, D.A., Rooney, A.A., Pickford, D.B., 1995.
Organization vs. activation: the role of endocrine-disrupting
chemicals (EDCS) during embryonic development in wildlife.
Environ. Health Perspect 103 (Suppl. 7), 157–164.
Giullette, L.J., Pickford, D.B., Crain, D.A., Rooney, A.A., Percival,
H.F., 1996. Reduction in penis size and plasma testosterone
concentrations in juvenile alligators living in a contaminated
environment. Gen. Comp. Endocrinol. 101, 32–42.
Goolish, E.M., Evans, R., Okutake, K., Max, R., 1998. Chamber
volume requirements for reproduction of the zebrafish (Brachydanio rerio). Prog. Fish Cult. 60, 127–132.
Gray, M.A., Metcalfe, C.D., 1997. Induction of testis-ova in Japanese
medaka (Oryzias latipes) exposed to 4-tert-octylphenol. Environ.
Toxicol. Chem. 16, 1082–1086.
Harries, J.E., Runnalls, T., Hill, E., Harris, C.A., Maddix, S.,
Sumpter, J.P., Tyler, C.R., 2000. Development of a reproductive
performance test for endocrine disrupting chemicals using pairbreeding fathead minnows (Pimephales promelas). Environ. Sci.
Technol. 34, 3003–3011.
Hood, T.E., Calabrese, E.J., Zuckerman, B.M., 2000. Detection of an
estrogen receptor in two nematode species and inhibition of
binding and development by environmental chemicals. Ecotoxicol.
Environ. Saf. 47, 74–81.
Huet, M.C., 2000. OECD activity on endocrine disrupters test
guidelines development. Ecotoxicology 9, 77–84.
Ingersoll, C.G., Hutchinson, T., Crane, M., Dodson, S., DeWitt, T.,
.
Gies, A., Huet, M.C., McKenney, C.L., Oberdorster,
E., Pascoe,
D., Versteeg, D.J., Warwick, O., 1999. Laboratory toxicity tests for
evaluating potential effects of endocrine-disrupting compounds. In:
deFur, P.L., Crane, M., Ingersoll, C., Tattersfield, L. (Eds.),
Endocrine Disruption in Invertebrates: Endocrinology, Testing and
Assessment. Society of Environmental Toxicology and Chemistry.
Pensacola, FL, USA, pp. 107–197.
Janssen, P.A.H., Lambert, J.G.D., Vethaak, A.D., Goos, H.J.T., 1997.
Environmental pollution causes elevated concentrations of estrogens and vitellogenin in the female flounder, Platichthys flesus.
Aquat. Toxicol. 39, 195–214.
Jobling, S., Nolan, M., Tyler, C.R., Brighty, G., Sumpter, J.P., 1998.
Widespread sexual disruption in wild fish. Environ. Sci. Technol.
32, 2498–2506.
Jobling, S., Sheahan, D., Osborne, J.A., Matthiessen, P., Sumpter,
J.P., 1996. Inhibition of testicular growth in rainbow trout
(Oncorhynchus mykiss) exposed to estrogenic alkylphenolic compounds. Environ. Toxicol. Chem. 15, 194–202.
Kavlock, R.J., Daston, G.P., deRosa, C., Fenner-Crisp, P., Earl Gray,
L., Kaatari, S., Lucier, G., Luster, M., Mac, M.J., Mazka, C.,
Miller, R., Moore, J., Rolland, G., Scott, G., Sheehan, D.M.,
Sinks, T., Tilson, H.A., 1996. Research needs for the risk
assessment of health and environmental effects of endocrine
disruptors. A report of the US EPA sponsored workshop. Environ.
Health Perspect 104, 715–740.
LaFont, R., 2000. The endocrinology of invertebrates. Ecotoxicology
9, 41–57.
Lye, C.M., Frid, C.L.J., Gill, M.E., McCormick, D., 1997. Abnormalities in the reproductive health of flounder, Platichthys flesus,
313
exposed to effluents from sewage treatment works. Mar. Pollut.
Bull. 34, 34–41.
Maack, G., Segner, H., 2002. Morphological development of the
gonads in the zebrafish Danio rerio. J. Fish Biol., in press.
Matthiessen, P., Waldock, R., Thain, J.E., Waite, M.E., 1995. Changes
in the periwinkle (Littorina littorea) populations following the ban
on TBT-based antifoulings on small boats in the United Kingdom.
Ecotoxicol. Environ. Saf. 30, 180–194.
Meregalli, G., Phuymers, L., Ollevier, F., 2001. Exposure of
Chironomus riparius larvae exposed to 4-n-nonylphenol. Environ.
Pollut. 111, 241–246.
Meregalli, G., Ollevier, F., 2001. Exposure of Chironomus riparius
larvae to 17a-ethinylestradiol: effects on survival and mouthpart
deformities. Sci. Total Environ. 269, 157–161.
Metcalfe, T.L., Metcalfe, C.D., Kiparissis, Y., Niimi, A.J., Foran,
C.M., Benson, W.H., 2000. Gonadal development and endocrine
response in the Japanese medaka (Oryzias latipes) exposed to o,p0 DDT in water or through maternal transfer. Environ. Toxicol.
Chem. 19, 1893–1900.
Miles-Richardson, S.R., Kramer, V.J., Fitzgerald, S.D., Render, J.A.,
Yamini, B., Barbee, S.J., Giesy, J.P., 1999. Effects of waterborne
expsoure to 17b-estradiol on secondary sex characteristics and
gonads of fathead minnows (Pimephales promelas). Aquat. Toxicol.
47, 129–145.
Nagel, R., 1993. Fish and environmental chemicals—a critical
evaluation of tests. In: Braunbeck, T., Hanke, W., Segner, H.
(Eds.), Fish Ecotoxicology and Ecophysiology. VCH, Weinheim,
pp. 174–178.
Nagel, R., Isberner, K., 1998. Testing of chemicals with fish—a critical
evaluation of tests with special regard to zebrafish. In: Braunbeck,
T., Hinton, D.E., Streit, B. (Eds.), Fish Ecotoxicology. Birkh.auser,
Basel, pp. 337–352.
Nimrod, A.C., Benson, B.H., 1998. Reproduction and development of
Japanese medaka following an early life stage exposure to
xenoestrogens. Aquat.Toxicol. 44, 141–156.
.
Oberdorster,
E., Cheek, A.O., 2000. Gender bender at the beach:
endocrine disruption in marine and estuarine organisms. Environ.
Toxicol. Chem. 20, 23–36.
OECD Organisation for Economic Cooperation and Development,
1992. OECD Guidelines for Testing of Chemicals: Fish, Early Life
Stage Toxicity Test, Section 2, Guideline 210. Paris, France.
OECD Organisation for Economic Cooperation and Development,
1993. OECD Guidelines for Testing of Chemicals: Fish, Prolonged
Toxicity Test: 14-Day-Study, Section 2, Guideline 204. Paris, France.
OECD Organisation for Economic Cooperation and Development,
1998. OECD Guidelines for Testing of Chemicals: Fish, Short-term
Toxicity Test on Embryo and Sac-fry Stages, Section 2, Guideline
212. Paris, France.
OECD Organisation for Economic Cooperation and Development,
1999. Fish Expert Consultation Meeting. London, Paris, France.
.
Orn,
S., Gessbo, A., Steinholz, A., Norrgren, L., 2000. Zebrafish
(Danio rerio—a candidate to evaluate endocrine disrupting
.
chemicals). In: Andersen, L., Bengtsson, B.E., Bjork,
M., et al.
(Eds.), Zebrafish for Testing Endocrine Disrupting Chemicals.
TemaNord 555, Copenhagen, pp. 47–62.
Pascoe, D., Carroll, K., Karntanut, W., Watts, M.M., 2002. Toxicity
of 17a-ethynylestradiol and bisphenol A to the freshwater
cnidarian Hydra vulgaris. Arch Environ. Contam. Toxicol. 43,
56–63.
Pinder, L.V.C., Pottinger, T.G., Billinghurst, Z., Depledge, M.H.,
1999. Endocrine function in aquatic invertebrates and evidence for
disruption by environmental pollutants. R&D Technical Report,
UK Environment Agency.
Purdom, C.E., Hardiman, P.A., Bye, V.J., Eno, N.C., Tyler, C.R.,
Sumpter, J.P., 1994. Estrogenic effects of the effluents from sewage
treatment works. Chem. Ecol. 8, 275–285.
314
H. Segner et al. / Ecotoxicology and Environmental Safety 54 (2003) 302–314
Roex, E.W.M., Giovannaneglo, M., van Gestel, A.M., 2001.
Reproductive impairment in the zebrafish, Danio rerio, upon
chronic exposure to 1,2,3-trichlorobenzene. Ecotoxicol. Environ.
Saf. 48, 196–201.
Sarojini, R., Jayalakshimi, K., Sambasivarao, S., 1986. Effect of
external steroids on ovarian development in freshwater prawn,
Macrobrachium lamerii. J. Adv. Zool. 7, 50–53.
Sarojini, R., Rao, S.S., Lakshmi, K.J., 1990. Effects of steroids
(estradiol and estrone) on the ovaries of the marine crab Scylla
serrata. Comp. Physiol. Ecol. 15, 21–26.
Scholz, S., Gutzeit, O., 2000. 17a-ethynylestradiol affects
reproduction, sexual differentiation and aromatase gene expression of the medaka (Oryzias latipes). Aquat. Toxicol. 50,
363–373.
Sharpe, R.M., Skakkebaek, N.E., 1993. Are oestrogens involved in
falling sperm counts and disorders of the male reproductive tract?
Lancet 341, 1392–1395.
Spooner, N., Gibbs, P.E., Bryan, G.W., Goad, L.J., 1991. The effect of
tributyltin upon steroid titres in the female dogwhelk, Nucella
lapillus, and the development of imposex. Mar. Environ. Res. 32,
37–49.
Sumpter, J.P., Jobling, S., 1995. Vitellogenesis as a biomarker for
estrogenic contamination of the aquatic environment. Environ.
Health Perspect. 103, 173–178.
Takahashi, H., 1977. Juvenile hermaphroditism in the zebrafish (Brachydanio rerio). Bull. Fac. Fish. Hokkaido Univ. 28,
57–65.
Tattersfield, L., Matthiessen, P., Campbell, P., Grandy, N., and L.ange,
R. (Eds.), 1997. SETAC Europ/OECD/EC Expert Workshop on
Endocrine Modulators and Wildlife: Assessment and Testing.
SETAC Europe, Brussels.
Tyler, C.R., Jobling, S., Sumpter, J.P., 1998. Endocrine disruption in
wildlife: a critical review of the evidence. Crit. Rev. Toxicol. 28,
319–361.
Vandenbergh, G.F., Adriaens, D., Janssen, C.R., 2002. Effects of 17aethinylestradiol on sexual development of the amphipod Hyalella
azteca. Ecotoxicol. Environ. Saf. 54 (2), 216–222.
Vos, J.G., Dybing, E., Greim, H.A., Ladefoged, O., Lambre, C.,
Tarazona, J.V., Brandt, I., Vethaak, A.D., 2000. Health effects of
endocrine-disrupting chemicals on wildlife, with special reference to
the European situation. Crit. Rev. Toxicol. 20, 71–133.
Wannemacher, R., Rebstock, A., Kulzer, E., Schrenk, D., Bock, K.W.,
1992. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin on reproduction and oogenesis in zebrafish (Brachydanio rerio). Chemosphere
24, 1361–1368.
Watts, M.M., Pascoe, D., Carroll, K.C., 2001a. Survival and
precopulatory behavior of Gammarus pulex (L.) exposed to
xenoestrogens. Water Res. 35, 2347–2352.
Watts, M.M., Pascoe, D., Carroll, K., 2001b. Chronic exposure to
xenoestrogens—effects on development and reproduction in the
freshwater invertebrate Chironomus riparius. Aquat. Toxicol. 55,
113–124.
Watts, M.M., Pascoe, D., Carroll, K.C., 2002a. Population responses of
the freshwater amphipod Gammarus pulex (L.) to an environmental
estrogen, 17a-ethinylestradiol. Environ. Toxicol. Chem. 21, 445–450.
Watts, M.M., Pascoe, D., Carroll, K., 2002b. Exposure to 17aethinylestradiol and bisphenol A—effects on larval moulting and
mouthpart structure of Chironomus riparius. Ecotoxicol. Environ.
Saf. 54 (2), 207–215.
Zou, E., Fingerman, M., 1997. Effect of estrogenic xenobiotics on
moulting of the water flea, Daphnia magna. Ecotoxical. Environ.
Saf. 38, 281–285.