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EFFECTS OF 17a-ETHYNYLESTRADIOL ON ENDOCRINE STATUS, REPRODUCTION, EARLY-LIFE DEVELOPMENT
AND SEXUAL DIFFERENTIATION IN MUMMICHOG (FUNDULUS HETEROCLITUS)
by
Rebecca Emily McLeod Peters
B.ScH, Queen’s University, 2000B.Com, Queen’s University, 2000
A Thesis Submitted in Partial Fulfillment of the Requirements for the Degree of
Master of Science
in the Graduate Academic Unit of Biology
Supervisor: Deborah MacLatchy, Ph.D., Biology
Examining Board: Karen Kidd, Ph.D., Biology (Chair) Kelly Munkittrick, Ph.D., Biology Lucy Wilson, Ph.D., Geology
This thesis, dissertation or report is accepted by the Dean of Graduate Studies
THE UNIVERSITY OF NEW BRUNSWICK
April, 2005
© Rebecca E. M. Peters, 2005
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Dedication
This thesis is dedicated to my grandmother, Sarah Agnes McLeod, who always believed
that education was the best gift a person could ever receive.
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Abstract
Adult mummichog (Fundulus heteroclitus) were exposed to 0, 0.1, 1, 10 or
lOOng/L 17a-ethynylestradiol (EE2) for 21 and 28 days in a reproductive bioassay
assessing endocrine and reproductive endpoints. The exposure protocol continued for 63
weeks with their offspring in an early life-stage bioassay assessing embryonic survival
and hatching, and a growout protocol assessing growth and reproductive potential.
Minimal disruption to adult reproductive endocrine status, offspring development and
reproductive potential occurred in fish exposed at environmentally-relevant
concentrations of EE2 (<10ng/L). In fish exposed to the highest concentration of EE2
tested (lOOng/L), vitellogenin was induced in adult male fish and juvenile females;
gonadosomatic index (GSI), spawning, fertilization, circulating steroid levels and gonadal
steroid production were affected in adult fish; and hatch success and male GSI decreased,
and the sex ratio was skewed to >80% female in the offspring. These effects have the
potential to affect long-term population sustainability. This work furthers our ability to
extrapolate between short-term endocrine response and population-level responses in
endocrine disruption studies.
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Preface
This thesis has been written in the articles format. Chapter 1 is a general
overview of the project, and the rationale behind its development. It outlines the main
research objectives, and how this study contributes to our understanding of endocrine
disruption. Chapter 2 describes a short-term reproductive bioassay, which assesses the
pre- and post-spawning endocrine status and reproductive potential of adult mummichog,
Fundulus heteroclitus, subsequent to a short-term exposure to 17a-ethynylestradiol (EE2).
Chapter 3 describes the effects of chronic EE2 exposure on the offspring of the adult fish
from Chapter 2, from embryo to maturity. Chapter 4 summarizes the results of the two
papers, discusses the combined significance of this research, and indicates future
directions for investigation.
Chapters 2 and 3 have been submitted to peer-reviewed journals, and have been
formatted according to the guidelines of these journals. Both Chapter 2 and 3 were
submitted under joint authorship with D. MacLatchy and S. Courtenay, who provided
valuable guidance and editorial assistance. The content of this thesis and the manuscripts
contained within were written by me.
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Acknowledgements
Though the name authoring this paper is mine, credit also lies with many others
who supported, challenged and assisted me throughout this process.
I want to thank my family for the amazing support that they have shown me. My
husband, Phillip Peters, an engineer who bravely accepted the challenge of proof-reading
a biology thesis, was amazingly supportive through the long weeks of endless water
changes, assays and writing. My daughter Olivia, who adapted well to long hours spent
at UNB in her early days, and long hours away from Mummy as she got older, forced me
to balance life and school. My step children, Reed and Julia, who loved to come and feed
the fish or give them a bath on the weekends, taught me to find the simple joy in science.
My parents, Margaret and Chris Ibey, my brothers, Andrew and Nathaniel, my sisters,
Jessica, and Rachael, and in-laws, Gerry Peters and Lorraine Day, helped with moving,
babysitting, and always had a sympathetic ear and words of support and encouragement.
Deborah MacLatchy was a fantastic supervisor. She always prefaced returning a
draft of this thesis with ‘It looks worse than it is©”, and it didn’t matter where in the
world she was, it was always back in my hands sooner than I wanted it to be! Her
feedback challenged me to become a better writer and to analyze things in ways I didn’t
know that I could. Her door was always open, be it a research question, dying fish,
balancing family and school, career choices or other problematic situations. Over the last
4 years, Deb not only became someone I respect as a scientist but a good friend.
My supervisory committee, Drs. Simon Courtenay and Tillmann Benfey, have
provided excellent feedback throughout the project. Drs. Kelly Munkittrick and Matt
Litvak have also given me good advice and suggestions.
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The MacLatchy and Munkittrick Labs (Rainie Sharpe, Lottie Vallis, Kevin
Shaughnessy, Chris Blanar, Genevieve Vallieres, Karen Gormley, Sandra Brasfield) and
other graduate students at UNBSJ (Dan Baker, Jennifer Peddle, Danny Jardine, Megan
Kirkpatrick, Roshini Kassie, Collin Arens) have been wonderful for donating their time
to long days of sampling tiny fish. Undergraduate students in the lab, in particular,
Jennifer Ings, Jennifer Adams and Leslie Carroll, were instrumental in the completion of
this project.
Barb Dowding and the Levy and Sirieix families gave me peace of mind by taking
such great care of Olivia when I was in class and at the lab. Wayne Armstrong, who
helped me build the new mummichog room, was always ready to lend me a tool or see
the bright side of the situation. Kelly Cummings, Alison McAslan, and Jo-anne Stevens
were always ready with advice, equipment and fumehood space.
This research was funded through a NSERC (National Sciences and Engineering
Research Council of Canada) Research Grant to Dr. MacLatchy as well as a Networks of
Centres of Excellence: Canadian Water Network grant to Dr. MacLatchy (PI: K.
Munkittrick).
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Dedication.................................................................................................................................ii
Abstract....................................................................................................................................iii
Preface...................................................................................................................................... iv
Acknowledgements..................................................................................................................v
Table of Contents...................................................................................................................vii
List of Tables........................................................................................................................... x
List of Figures.... ................................................................................................................... xii
List of Abbreviations............................................................................................................xvi
Chapter 1: General Introduction........................................................................................1Development of a life-cycle bioassay for mummichog (Fundulus heteroclitus) for use in endocrine disruption studies
Introduction...............................................................................................................................2
Sewage Treatment Plants and 17a-Ethynylestradiol............................................................ 3
Endocrine Disruption...............................................................................................................6
Reproduction in Fish................................................................................................................7
Short-Term Endocrine Studies..............................................................................................10
Lifecycle Bioassays for Endocrine Disruption Studies......................................................12
Fish Species........................................................................................................................... 13
Study Objectives....................................................................................................................16
References...............................................................................................................................17
Chapter 2*: Endocrine Status and Reproduction.......................................................... 29Effects on reproductive potential and endocrine status in the mummichog {Fundulus heteroclitus) after exposure to 17a-ethynylestradiol in a short-term reproductive bioassay
Abstract 30
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Introduction 31
Materials and Methods.......................................................................................................... 35
Chemicals....................................................................................................................... 35
Fish..................................................................................................................................35
Exposures....................................................................................................................... 36
Fecundity measures....................................................................................................... 37
Fish sampling and reproductive endocrine endpoints.................................................39
Radioimmunoassay for steroids................................................................................... 40
Vitellogenin assay......................................................................................................... 41
Statistics......................................................................................................................... 41
Results.....................................................................................................................................42
21-day exposure.............................................................................................................42
28-day exposure.............................................................................................................44
Fecundity Measurements..............................................................................................53
Discussion...............................................................................................................................56
Acknowledgements................................................................................................................65
References.............................................................................................................................. 66
Chapter 3*: Early-life Development and Sex Differentiation...................................... 75Effects of 17a-ethynylestradiol on early-life development, sex differentiation and vitellogenin induction in mummichog (Fundulus heteroclitus)
Abstract...................................................................................................................................76
Introduction.............................................................................................................................77
Materials and Methods 79
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Chemicals................................................................................................................................79
Exposures................................................................................................................................79
Exposures - offspring.............................................................................................................80
Statistics..................................................................................................................................84
Results.....................................................................................................................................85
Discussion...............................................................................................................................97
Acknowledgements..............................................................................................................109
References............................................................................................................................ 110
Chapter 4: General Discussion........................................................................................118The effects of chronic exposure to 17a-ethynylestradiol on the lifecycle of mummichog (Fundulus heteroclitus)
References............................................................................................................................ 126
Curriculum Vitae
* Chapters 2 and 3 have been submitted for external review and publication.
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Table 2.1 43
Mean [±1SE] weight, length, liver somatic index (LSI), gonadosomatic index
(GSI) and condition factor for adult male and female Fundulus heteroclitus
exposed to EE2 for 21 and 28 days during gonadal recrudescence (21 days) and
spawning (28 days) in June 2003.
Summary of changes in plasma vitellogenin, plasma titres and gonadal production
of testosterone (T), 11-ketotestosterone (11-KT; males only) and 17P-estradiol
(E2; females only) and spawning (females only) and fertilization (males only) in
mummichog exposed to graded doses of waterborne ethynylestradiol (EE2) for 21
or 28 days.
Mean [±1SE] weight, length, gonadosomatic index (GSI), Liver somatic index
(LSI) and condition factor for juvenile male and female Fundulus heteroclitus
exposed to 17a-ethynylestradiol (EE2) from fertilization until 48, 52 and 61 weeks
after hatch.
Table 2.2 57
Table 3.1 91/92
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Table 3.2..................................................................................................................................93
p-values for changes in morphological variables of juvenile mummichog among
weeks 48, 52, and 61 post-hatch after exposure to 0, 0.1, 1, 10 or lOOng/L 17a-
ethynylestradiol (EE2).
Table 3.3............................................................................................................................... 100
Proportion (mean ± 1SE) of juvenile mummichog exhibiting female secondary
sex characteristics and/or female gonads at 48, 52, and 61 weeks post-hatch after
exposure to 0, 0.1, 1, 10 or lOOng/L EE2.
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Figure 1.1................................................................................................................................. 8
The reproductive system in fish, including the hypothalamo-pituitary-gonadal
axis. Modified from Kime 1998, Kime 1999 and McMaster et al., 1995. (GnRH
= gonadotropin releasing hormone; GtH = gonadotropin; T = testosterone; E2 =
estradiol; 11KT = 11-ketotestosterone; + and - represent positive and negative
feedback loops).
Figure 2.1................................................................................................................................45
Plasma vitellogenin (VTG) levels in male and female mummichog exposed to
0, 0.1, 1, 10 or lOOng/L 17a-ethynylestradiol (EE2) for 21 (A) or 28 (B) days.
Figure 2.2 (A&C)....................................................................................................................46
Plasma testosterone (T) and 11-ketotestosterone (11KT) from male
mummichog exposed to 0, 0.1,1,10 or lOOng/L 17a-ethynylestradiol (EE2)
during gonadal recrudesence (2Id - A) or during spawning (28d - C).
Figure 2.2 (B&D).................................................................................................................... 47
Plasma T and 17p-estradiol (E2) from female mummichog exposed to 0, 0.1,1,
10, lOOng/L EE2 during gonadal recrudesence (2Id - B) or spawning (28d - D).
Figure 2.3........................................................................................................................... 48/49
(A) In vitro testosterone production in M l99 plus 3-isobutyl 1-
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methylxanthine (IBMX; ImM) (basal) and in M l99 plus IBMX and human
chorionic gonadotropin (20 IU/mL; hCG) - enhanced gonadal incubations from
male fish exposed to 0, 0.1,1,10, or 100ng/L 17a-ethynylestradiol (EE2) for 2Id.
(B) In vitro testosterone production in M l99 plus IBMX (basal) and M l99 plus
IBMX and hCG-enhanced gonadal incubations from female fish exposed to 0,
0.1, 1, 10 or 100ng/L EE2 for 2 Id. (C) In vitro estradiol production in M l99 plus
IBMX (basal) and in M l99 plus IBMX and hCG - enhanced gonadal incubations
from female fish exposed to 0, 0.1, 1, 10, or 100ng/L EE2 for 2 Id.
Figure 2.4........................................................................................................................... 51/52
(A) In vitro testosterone production in M l99 plus 3-isobutyl 1-
methylxanthine (IBMX; ImM) (basal) and in M l99 plus IBMX and human
chorionic gonadotropin (20 IU/mL; hCG) - enhanced gonadal incubations from
male fish exposed to 0, 0.1,1, 10, or 100ng/L 17a-ethnylestradiol (EE2) for 28d.
(B) In vitro testosterone production in M l99 plus IBMX (basal) and M l99 plus
IBMX and hCG-enhanced gonadal incubations from female fish exposed to 0,
0 .1,1,10 or 100 ng/L EE2 for 28d. (C) In vitro estradiol production in M199 plus
IBMX (basal) and in M l99 plus IBMX and hCG (hCG) - enhanced gonadal
incubations from female fish exposed to 0, 0.1,1, 10, or 100ng/L EE2 for 28d.
Figure 2.5.................................................................................................................................54
Mean cumulative eggs spawned per female mummichog exposed to 0, 0.1, 1, 10
or 100ng/L 17a-ethynylestradiol.
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Figure 2.6.................................................................................................................................55
Mean fertilization rate of mummichog eggs calculated daily between days 21
and 28 of a 28-day exposure to 0, 0.1, 1, 10 or lOOng/L 17a-ethynylestradiol.
Figure 3.1........................................................................................................................... 86/87
(A) Mean time to hatch of mummichog eggs exposed to 0, 0.1, 1, 10, or lOOng/L
17a-ethynylestradiol (EE2). (B) Mean hatch success of mummichog eggs
exposed to 0, 0.1, 1, 10, or lOOng/L EE2 . (C) Mean length at hatch of
mummichog larvae exposed to 0, 0.1, 1, 10, or lOOng/L EE2 .
Figure 3.2........................................................................................................................... 89/90
(A) von Bertalanffy growth curves for mummichog larvae exposed to 0, 0.1, 1,
10, or lOOng/L 17a-ethynylestradiol. (B) Mean total length of mummichog
exposed to 0, 0.1, 1, 10, or lOOng/L 17a-ethynylestradiol between week 1 and
week 63 post-hatch.
Figure 3.3.................................................................................................................................94
Mean survival of mummichog exposed to 0, 0.1, 1, 10, or lOOng/L 17a-
ethynylestradiol from hatch to 61 weeks post-hatch.
Figure 3.4.................................................................................................................................96
Mean prevalence of vertebral abnormalities in juvenile mummichog exposed to 0,
0.1, 1, 10, or lOOng/L 17a-ethynylestradiol at 15, 48 and 61 weeks post-hatch.
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Figure 3.5.................................................................................................................................98
Liver vitellogenin levels in male and female mummichog exposed to 0, 0.1,1,10
or lOOng/L 17a-ethynylestradiol (EE2) for 52 weeks post-hatch.
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ANCOVA Analysis of CovarianceANOVA Analysis of VariancecAMP 3’,5’-cyclic adenosine monophosphateCF Condition FactorCl Confidence IntervalDO Dissolved OxygenEi Estronee 2 17ß-estradiolEC50 Median Effective ConcentrationEDS Endocrine Disrupting Substancee e 2 17 a-ethynylestradiolELISA Enzyme-Linked Immunosorbant AssayEtOH EthanolGnRH Gonadotropin Releasing HormoneGSI Gonadosomatic IndexGtH Gonadotropin HormonehCG Human chorionic gonadotropinIBMX 3-isobutyl 1-methylxanthineIU International UnitsLSI Liversomatic IndexM199 Medium 199RIA RadioimmunoassaySE Standard ErrorSTP Sewage Treatment PlantT TestosteroneVTG Vitellogenin11-KT 11-ketotestosterone
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Chapter 1: General Introduction
Development of a life-cycle bioassay for mummichog (Fundulus
heteroclitus) for use in endocrine disruption studies
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Introduction
Organisms that occupy aquatic habitats are at great risk of contaminant
exposure, as chemicals released into air, land or water can eventually be
transported to the water (Kime, 1999). Fish living downstream of municipal
(Harries et al., 1996; Jobling et al., 1998; Larsson et al., 1999) and industrial
(LeBlanc et al., 1997; Black et al., 1998) effluents have demonstrated impacts on
reproductive potential, including the presence of intersex, delayed reproduction,
and vitellogenin (VTG) induction in male fish. Some of these effects have been
linked to natural and synthetic steroidal estrogens identified in municipal waste
water from sewage treatment plants (STPs) (Routledge et al., 1998). Short-term
exposure to one of these steroids, 17a-ethynylestradiol (EE2), induced VTG and
altered sex steroid production and circulating levels in mummichog (Fundulus
heteroclitus) (MacLatchy et al., 2003), indicating possible impacts on
reproduction.
The objectives of this research project were to develop a short-term
reproductive bioassay in mummichog to see if reproductive endocrine endpoints
in adult fish could be associated with reproductive potential and to determine the
impact of chronic EE2 exposure on offspring development. By linking these
objectives, I wished to determine whether effects caused by endocrine disrupting
substance (EDS) exposure seen in short-term bioassays may be linked to
reproductive potential and progeny impacts. These effects may translate to
population effects when fish are exposed to EE2 throughout their entire lifecycle.
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Sewage Treatment Plants and 17a-Ethynylestradiol
For centuries, aquatic environments have been receiving municipal and
industrial wastes which contain thousands of natural and synthetic chemicals.
Population increases, along with technological and societal advances, are
demanding treatment and recycling of available water resources. Sewage
treatment plants (STPs) improve the quality of municipal water released back to
the aquatic environment. Processes differ among STPs, resulting in variable
performance overall (Routledge et al., 1998).
STP effluents are tested on aquatic organisms to determine safe discharge
levels. However, this testing does not necessarily include toxicity testing on
individual chemical components of the effluent and/or their hormone-disrupting
ability (Desbrow et al., 1998). Studies have shown that both wild and caged fish
have significant estrogen responses downstream of STP outfalls, including the
presence of ovarian tissue in the testes, impairment of gonadal development,
alterations in sex steroid levels, and VTG induction in male fish (Harries et al.,
1996; Jobling et al., 1998; Larsson et al., 1999; Folmar et al., 2000). Estrogen
receptor agonists (estrogenic compounds) have been identified in STP effluents
and include alkylphenols and natural and synthetic estrogens (Desbrow et al.,
1998; Temes et al., 1999a; Nasu et al., 2001).
Estrogens are not completely removed during sewage treatment. Those
most commonly detected in STP effluents are estrone (Ei) and 17p-estradiol (E2),
estrogens naturally produced and excreted, and 17a-ethynylestradiol (EE2), a
synthetic hormone (Desbrow et al., 1998; Temes et al., 1999a; Snyder et al.,
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2001). The loads entering the aquatic environment are relatively small (lower
ng/L range), however, when combined, these hormones can show estrogenic
effects in fish at individual steroid no-effect levels, indicating that the total
estrogenic load of STP effluent may be more potent than its individual
components (Thorpe et al., 2003). The potency of EE2 (as assessed by the
median effective concentration (EC50) of vitellogenin response in juvenile
rainbow trout) is as much as 27 times E2 and E2 is more potent than Ei by up to
3.2 times (Thorpe et al., 2003). EE2 is a pharmaceutical compound, the main
estrogenic ingredient of the combined birth control pill, which typically contains
30-50jj,g EE2 per pill (Desbrow et al., 1998). In Canada, concentrations of EE2
typically fall in the range of l-10ng/L in STPs although levels have been
documented as high as 42ng/L (Desbrow et al., 1998; Ternes et al., 1999a), and
downstream of the effluent discharge estrogen load is further diluted by river
water and rain water (Harries et al., 1996). Unlike the natural estrogens, EE2 is
highly stable in activated sludge and is not broken down in sewage treatment
(Ternes et al., 1999b), making it a persistent and environmentally-relevant
compound. For these reasons, EE2 has been listed as a reference chemical for
EDS testing by the Organization for Economic Cooperation and Development
(OECD, 1999).
Methods used to expose fish to EE2 have included injection into the
developing egg, contamination of feed, and waterborne exposures (Blâzquez et
al., 1998; Papoulias et al., 1999; MacLatchy et al., 2003). Though some studies
have demonstrated non-reproductive impacts on fish, such as alterations in the
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kidney, liver and spleen of common carp (Cyprinus carpio) (Schwaiger et al.,
2000) and kidney and liver cell death in medaka (Oryzias latipes) (Weber et al.,
2004), most studies focus on reproductive impacts. Complete sex-reversal of
genetically male medaka has been shown following injection of 0.5-5ng/egg EE2
directly into the oil globule of fertilized eggs (Papoulias et al., 1999). Water
borne exposure of medaka to EE2 post-hatch (until sexual maturity) results in
increased incidence of intersex and impaired reproductive behaviour in males,
indicating that EE2 may not prohibit reproduction, but reduces male reproductive
performance (Balch et al., 2004). Furthermore, quantification of gonadal cell
death after chronic exposure to EE2 suggests that gonadal toxicity is restricted to
male fish in medaka (Weber et al., 2004).
Sea bass (Dicentrarchus labrax) were exposed to EE2 in food during a
developmental period previously defined as sensitive to androgen exposure.
Upon maturation, the number of female fish doubled compared to controls, and
those fish not identified as female had suppressed gonadal development.
However, estrogenic exposure after this sensitive period had no impact on the sex
ratio, though composition of male reproductive organs may have been altered
(Bläzquez et al., 1998).
Studies with fathead minnow (Pimephales promelas) demonstrated that
male fish failed to develop secondary sex characteristics and gonadal sex ratios
were skewed at 4ng/L EE2 after 56 days and no testis tissue was found at 172 days
post-hatch (Länge et al., 2001). After a 24h exposure to concentrations as low as
2ng/L, adult male and 48h post-hatch larval fathead minnow were transcribing
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vitellogenin (YTG) gene products (Lattier et al., 2002). Short-term adult EE2
exposures with mummichog (Fundulus heteroclitus) have found concentration-
dependent VTG induction in addition to altered circulating steroid levels and
steroidogenesis in both sexes (MacLatchy et al., 2003).
Endocrine Disruption
Conventional toxicity tests assess exogenous pollutants in terms of
mortality, stress, discomfort, growth, development and illness in the exposed
organism. However, though dead fish clearly indicate that a compound or
mixture is toxic, seemingly healthy fish may in fact be masking a threat to the
population. At levels of exposure well below those that elicit stress response or
mortality, reproductive dysfunction can occur.
Bodily functions such as fluid homeostasis, stress management, and
reproduction are regulated by the endocrine system (Kime, 1998). This system is
a chemical communication network comprised of organs such as the gonads,
pituitary, hypothalamus, thyroid and liver that synthesizes and circulates
hormones that regulate processes such as growth, development and reproduction.
This system is extremely sensitive to low levels of pollutants and action by
exogenous substances at any tissue involved in control of reproduction
(hypothalamus, pituitary, gonad, liver or gametes) can have an impact on gamete
quantity or quality resulting in reproductive dysfunction. Action on the endocrine
system can occur through interference with the production, release, transport,
metabolism, binding, action or elimination of natural hormones (Kavlock et al.,
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1996). True disruption of the endocrine system occurs when, subsequent to action
on the endocrine system by an exogenous substance, there is an effect on the
health of the intact organism or its progeny (OECD, 1999). Without the whole
organism or progeny effect, a compound capable of altering the hormonal balance
is deemed to be hormonally-active rather than an EDS.
Of greatest functional significance in EDS studies are the impacts on
reproductive success (Ankley et al., 2001; Grist et al., 2003). Effects of exposure
to potential EDSs may be manifested in poor quality of gametes (Kime and Nash,
1999), altered gonadosomatic index (GSI; LeBlanc et al., 1997), decreased egg
production or fertilization (Kime, 1998), altered sex ratios and/or intersex (Jobling
et al., 1998; Parrott and Wood, 2002), altered behaviour that impedes mating
(Balch et al., 2004), larval survival and development and future fecundity (Kime,
1998). These parameters imply a general reduction in reproductive success;
however, differences in sensitivities exist among species (Grist et al., 2003) and
life stages (Boudreau et al., 2004) which will affect the degree to which the
population is impacted.
Reproduction in Fish
A species-specific set of external cues (e.g., temperature, photoperiod,
lunar and tidal cycles) control gonadal recrudescence in fish through the
hypothalamo-pituitary-gonadal axis (Fig. 1.1; Kime, 1999). In teleost fish, the
hypothalamus releases gonadotropin releasing hormone (GnRH) which stimulates
gonadotropin (GtH) production in the pituitary. Gonadotropins act upon the
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External Cues Internal Biological Regulation
Figure 1.1. The reproductive system in fish, including the hypothalamo-pituitary-
gonadal axis. Modified from McMaster et al., 1995, Kime 1998, and Kime 1999.
(GnRH = gonadotropin releasing hormone; GtH = gonadotropin; T = testosterone;
E2 = estradiol; 11KT = 11-ketotestosterone; + and - represent positive and
negative feedback loops).
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gonad to induce gonadal growth, gamete development, sex steroid production and
gamete maturation (Kime, 1999).
The hypothalamo-pituitary-gonadal axis is controlled by positive and
negative feedback loops from androgens and estrogens produced in the gonad.
These sex steroids control the synthesis and release of gonadotropins to regulate
growth and maintenance of reproductive organs, gamete production and sexual
behaviour (Kime, 1999). Testosterone is produced in male and female fish and
serves as a precursor to 11-ketotestosterone in males and 17p-estradiol in females.
11-Ketotestosterone regulates spawning behaviour, secondary sex characteristics
expression and sperm cell maturation in male fish (Cochran et al., 1988). In
females, 17p-estradiol stimulates hepatic production of vitellogenin, a protein
precursor to egg yolk, often used as a biomarker for estrogen exposure (Kime,
1999).
Internal homeostasis must be maintained to ensure reproductive success.
Uptake of hormonally-active substances and/or EDSs that can mimic natural
hormones, inhibit hormonal action and/or interfere with hormone receptors to
alter the normal feedback loops to the pituitary and hypothalamus may result in
altered GtH secretion (Kime, 1999). Gonadotropin initiates steroidogenesis, and
through gonadal steroids, is ultimately responsible for secondary sex
characteristics, courtship behaviour, gamete production and maturation, and
spawning; thus, interruption of pathways controlling GtH release by EDSs can
affect normal reproduction.
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Estrogenic alkylphenols can displace gonadal E2, bind at multiple estrogen
receptor sites, and alter the conformation of the receptor complex affecting
regular feedback within the gonad (Mueller and Kim, 1978). The sex steroid
biosynthetic pathway, which converts cholesterol to estradiol or testosterone
through a series of intermediaries, is controlled by regulatory enzymes, which
may be inhibited or stimulated by exposure to EDSs or hormonally-active
substances (McMaster et al., 1995; Conde^a and Canario, 1999). Hormonally-
active substances and/or EDSs can also alter gonadal composition and/or
secondary sex characters to feminize estrogen-exposed male fish (Blazquez et al.,
1998; Conde^a and Canario, 1999) and masculinize androgen-exposed female fish
(Gale et al., 1999; Ankley et al., 2001).
The changes to the reproductive system from alterations of the
hypothalamic-pituitary-axis may manifest as affected reproductive potential (e.g.
hormone synthesis, reproductive behaviours) and/or population viability (e.g.
fecundity, fertilization, sex ratios). Changes to reproductive potential and
population viability from EDS exposure have been assessed in short- and long
term laboratory studies (Harries et al., 2000; Ankley et al., 2001; MacLatchy et
al., 2003; Schultz et al., 2003; Seki et al., 2003).
Short-Term Endocrine Studies
As negative consequences of the impacts of chemical exposure in fish
continue to be revealed, an international movement to develop standardized
laboratory fish bioassays to examine the impacts of EDSs and hormonally-active
Page 28
substances has arisen (Gray et al., 1997; OECD, 1999; Ankley et al., 2001).
Short-term laboratory tests may be done with adult fish (Ankley et al., 2001;
MacLatchy et al., 2003) or during early life stages (Couillard, 2002; Boudreau et
al., 2004).
For EDS testing purposes, short-term laboratory tests are effective, as
results may be produced relatively quickly and cost effectively, in a controlled
setting. Partial and short-term bioassays for EDS exposure have been developed
for freshwater fish such as fathead minnow, Pimephalespromelas (Harries et al.,
2000; Lattier et al., 2002), rainbow trout, Oncorhynchus mykiss (Schultz et al.,
2003), and medaka, Oryzias latipes (Seki et al., 2003) and for estuarine and
marine species such as sheepshead minnow, Cyprinodon variegates (Folmar et
al., 2000; Zillioux et al., 2001) and mummichog, Fundulus heteroclitus
(MacLatchy et al., 2003; Boudreau et al., 2004). However, the endpoints
measured in each study differ and range from induction of VTG to endocrine
status (e.g. hormone synthesis and circulating profiles) to basic reproductive
endpoints (e.g. gonad size, secondary sex characteristics, fecundity) to
developmental attributes (e.g. abnormalities, mortality, growth). In order to better
diagnose effects and/or extrapolate across species, bioassays must consistently
assess endpoints that reflect typical actions of the classes of EDS. To project
population-level impacts of EDSs, bioassays must also assess basic yet critical
aspects of reproduction and early life history (Ankley et al., 2001). Assessing
these endpoints will help us to better understand the impacts of EDSs and/or
hormonally-active compounds and mixtures.
Page 29
Lifecycle Bioassays for Endocrine Disruption Studies
Short-term bioassays are essential tools for the screening and assessment
of EDSs and hormonally-active compounds and mixtures in a timely and cost-
effective manner. However, endpoints measured in short-term exposures provide
only a snapshot of the total effect(s) of exposure and may actually show impacts
on the endocrine system, VTG or other biochemical parameters that are
eliminated by homeostatic mechanisms during chronic exposure. Furthermore,
chronic exposure can induce changes to the organism not noted by shorter-term
exposures, including factors that may be manifested at the population level, such
as lower recruitment (Ankley et al., 2001). Short-term tests may also have
lowered sensitivity compared to tests that assess impacts at more than one life
stage (Parrott and Wood, 2002). Finally, endocrine disruption requires an effect
on the health of an intact organism or its progeny subsequent to exposure and
alteration of endocrine status; lifecycle testing addresses whether exposure of
parental fish has adverse effects on the next generation (Seki et al., 2003). It is
difficult to assess which endpoints are most sensitive to EDS exposure. Many
factors including life stage of the fish and mechanism-related (e.g. VTG, sex
steroids) and reproductive (e.g. secondary sex characteristics, reproductive output,
behaviour, gonadal histology) endpoints must be investigated in lifecycle tests in
order to develop short-term tests that are sensitive and predictive of long-term,
chronic and multigenerational impacts (Parrott and Wood, 2002).
Page 30
Responses of saltwater species to EDSs, in both short and long-term
laboratory exposures, have been overlooked compared to freshwater responses,
perhaps because fewer standardized toxicity tests exist for saltwater species and
aquatic risk assessments have focused more on freshwater species (Leung et al.,
2001). The effect of salinity on bioavailability of exogenous chemicals has not
been well studied (El-Alfy, et al., 2002) though increased copper uptake was
found in mummichog exposed to high salinity (Blanchard and Grosell, 2003) and
increased salinity enhanced the toxicity of pesticides in medaka (El-Alfy and
Schlenk, 1998). Long and short-term bioassays for estuarine fish have been
developed with sheepshead minnow, Cyprinodon variegates (Manning et al.,
1999; Folmar et al., 2000; Zillioux et al., 2001; Karels et al., 2003) and
mummichog, (Matta et al., 2001; Urushitani et al., 2002; MacLatchy et al., 2003;
Boudreau et al., 2004; Sharpe et al., 2004). This study is the first with
mummichog to assess reproductive and indicator endpoints throughout the
lifecycle to determine long-term impact of endocrine disruption observed in short
term EDS adult bioassays (MacLatchy et al., 2003; Sharpe et al., 2004).
Fish Species
The mummichog (Fundulus heteroclitus) is widely distributed along the
Eastern seaboard of North America, with a geographical distribution from the
Gulf of St. Lawrence to Texas (Armstrong and Child, 1965; Lotrich, 1975; Scott
and Scott, 1988). They are abundant across that range, and are easily collected
with beach seines or minnow traps (Eisler, 1986). Mummichog inhabit shallow,
Page 31
brackish water coves, inlets and tidal creeks that are regularly subjected to large
changes in salinity and water level (Lotrich, 1975). The mummichog has adapted
to reside in a rapidly-changing tidal environment and is tolerant to wide
fluctuations in dissolved oxygen, pH, temperature (3°C to 45°C) and salinity
(0.0%o to 120.3%o) (Eisler, 1986; Scott and Scott, 1988). Much is known about
the biology of this species, as it has been used in embryological, ecological,
physiological, molecular and environmental studies (Armstrong and Child, 1965;
Oppenheimer, 1979; Eisler, 1986; Taylor, 1986).
Their adaptability to differing environmental conditions and small size
(rarely exceeding 90mm; Scott and Scott, 1988), make mummichog suitable for
use in laboratory (MacLatchy et al., 2003; Sharpe et al., 2004), artificial stream
(Dube et al., 2002), and field (Leblanc et al., 1997; Couillard and Nellis, 1999)
studies. Mummichog have demonstrated sensitivity to xenobiotics such as
insecticides (Eisler, 1986), pulp mill effluents (LeBlanc et al., 1997; Dube et al.,
2002), PCBs (Black et al., 1998), and model endocrine disrupting compounds
(MacLatchy et al., 2003; Sharpe et al., 2004). Laboratory studies have shown that
mummichog sensitivity to hormonally-active substances may change with
different lifecycle stages, adults being more sensitive (MacLatchy et al., 2003;
Sharpe et al., 2004) than embryos/larvae (Matta et al., 2001; Urushitani et al.,
2002; Boudreau et al., 2004).
Sex ratios may be easily controlled in laboratory studies, as adult
mummichog are sexually dimorphic. Male fish are dark dorsally (blue-black or
dark green) with a black dot on the dorsal fin; belly, pelvic and caudal fin are
Page 32
brilliant yellow during the reproductive season, with pale, silvery vertical stripes
on the trunk and tail (Armstrong and Child, 1965; Scott and Scott, 1988). Female
fish have blackish-green backs, grading to white bellies, possibly with faint dark
striping on the tail (Armstrong and Child, 1965; Scott and Scott, 1988). These
colors are more evident during the breeding season.
The breeding season in mummichog begins in the spring (March-May)
and ends in the late summer or early autumn (July-September), and is longer in
the southern end of its range (Taylor et al., 1979; Hsiao et al., 1994).
Environmental cues (photoperiod and temperature) induce gonadal recrudescence
and regulate subsequent spawning events via lunar phases and tidal cycles.
Spawning events occur with the new and/or full moon (Taylor et al., 1979; Day
and Taylor, 1984; Taylor, 1986). Manipulation of photoperiod and temperature
can induce spawning in the laboratory in off-season (MacLatchy et al., 2003;
MacLatchy et al., 2005) providing a year-round supply of recrudescing fish.
In the wild, female mummichog use multiple substrates for spawning,
including empty mussel shells, marsh grasses (Spartina altemiflora), algae mats,
sand, and plant roots (Taylor et al., 1979). In the laboratory, mummichog will
spawn on algae mats (Couillard, 2002), against a mesh cage inside the aquarium
(Urushitani et al., 2002) or may be stripped and fertilized manually (Kelly and Di
Guilio, 2000; Matta et al., 2001). Embryos are hardy, able to withstand
desiccation, temperature and salinity changes and low oxygen levels (Armstrong
and Child, 1965; Taylor et al., 1979). Furthermore, eggs are a good size for
operative procedures (2.0mm) (Armstrong and Child, 1965; Taylor, 1984) and
Page 33
have a clear chorion, allowing views of embryonic development (Armstrong and
Child, 1965). Sexual maturation in mummichog occurs after the fish has reached
38mm (female) and 25mm (male), which may be reached in as little as six months
in some populations (Kneib and Stiven, 1978; Taylor et al., 1979). These
characteristics make mummichog very suitable for lifecycle bioassay
development.
Study Objectives
This research was undertaken to determine the sensitive life stages (if any)
and potential population-level impacts of chronic estrogen exposure on
mummichog. A lifecycle bioassay, consisting of a short-term adult reproductive
bioassay, an early life-stage bioassay and a growout protocol, were developed for
mummichog. It is important to link adult endocrine system status after short-term
exposure and measures of adult reproductive potential and/or development,
survival and reproductive potential of the offspring. This would further our
ability to extrapolate between short-term endocrine responses and population-
level responses and begin to address the gap in the literature regarding the impact
of EDS exposure during the life cycle of saltwater or estuarine fish.
Page 34
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Page 46
Chapter 2*: Endocrine Status and Reproduction
Effects on reproductive potential and endocrine status in the
mummichog (Fundulus heteroclitus) after exposure to 17a-
ethynylestradiol in a short-term reproductive bioassay
*This chapter was submitted to the journal Aquatic Toxicology under joint authorship
with Simon C. Courtenay (Fisheries and Oceans Canada) and Deborah L. MacLatchy
(University of New Brunswick)
Page 47
Abstract
A short-term reproductive bioassay with the mummichog (Fundulus heteroclitus)
was developed to link changes in endocrine status to reproductive potential subsequent to
endocrine disrupting substance (EDS) exposure. Sexually-mature mummichog were
separated by sex and exposed to the synthetic estrogen 17a-ethynylestradiol (EE2) at
concentrations of 0 to lOOng/L for 21 days using a static daily renewal protocol. Half of
the fish were sampled on Day 21. At lOOng/L, male fish had induction of vitellogenin
(VTG), increased gonadosomatic index (GSI), decreased testosterone production and
decreased circulating 11-ketotestosterone (11-KT). Female fish had decreased circulating
estradiol (E2) and testosterone (T) at lOOng/L. There were some impacts at lower
concentrations of EE2 in both sexes, though the results were not consistent. On Day 21,
the remaining male and female fish were combined at each treatment (three of each
sex/tank). These fish were exposed for an additional seven days during which spawning
and fertilization success were assessed in addition to the endocrine parameters measured
on Day 21. Males exposed to lOOng/L EE2 exhibited VTG induction, increased GSI, and
decreased testosterone production on Day 28. Female fish had increased E2 and T
production at 1 and lOng/L and circulating E2 levels remained depressed above lOng/L.
Female fish exposed to lOOng/L spawned fewer eggs than all other treatments;
fertilization was also impaired. Fish exposed to concentrations of EE2 below
environmentally relevant levels (i.e., <10ng/L) showed minimal effects on reproductive
status while both the endocrine system and reproductive potential were affected at
lOOng/L EE2.
Page 48
Key Words
Ethynylestradiol, Fundulus heteroclitus, vitellogenin, reproductive steroids, reproduction,
endocrine disruption
Introduction
Laboratory and wild fish studies have shown that exposure to endocrine
disrupting substances (EDSs), mixtures or effluents can have an impact on the
endocrinology of adult fish (Routledge et al., 1998; Dube and MacLatchy, 2001; Sharpe
et al., 2004). Although short-term adult fish bioassays are a tool for identifying
compounds with the potential to affect fish endocrine status (Kovacs et al., 1995),
protocols must be further developed into multigenerational studies to identify the long
term reproductive impacts, if any, due to exposure (Ankley et al., 2001).
Transgenerational studies have shown impacts on adult fish (survival, morphology,
histology, fecundity) and their offspring (hatching, growth, survival, sex ratios) (Patyna
et al., 1999; Matta et al., 2001; Zillioux et al., 2001; Parrott and Wood, 2002), however,
few studies have attempted to experimentally link physiological changes in adult fish
with individual or population-level fitness (Ankley et al., 2001).
Resource demands, and the time required to complete a life cycle, make longer-
term bioassays challenging. However, life-cycle bioassays provide information not
available from short-term exposures. Life-cycle bioassays have been developed for many
freshwater fish including the fathead minnow, Pimephalespromelas (Ankley et al., 2001;
Page 49
Länge et al., 2001), Japanese medaka, Oryzias latipes (Patyna et al., 1999) and zebrafish,
Danio rerio (Olsson et al., 1999). Estuarine and marine species have been studied to a
lesser degree, although short-term bioassays have been developed with sheepshead
minnow, Cyprinodon variegates (Folmar et al., 2000; Zillioux et al., 2001) and more
recently mummichog, Fundulus heteroclitus (Urushitani et al., 2002; MacLatchy et al.,
2003; Boudreau et al., 2004; Sharpe et al., 2004).
The mummichog, an estuarine killifish, is a good candidate for bioassay
development. Breeding is cyclical where gonad size, measured as a percentage of total
body weight, and steroid hormone levels coincide with spawning events on the full and
new moons during the summer months in field and controlled laboratory conditions
(Taylor et al., 1979; Cochran, 1987; Hsiao and Meier, 1989). A preliminary study
showed that fish captured in New Brunswick had a large spawning event at the full moon
and a smaller spawning event at the new moon (unpublished data). Gonadal
recrudescence and natural spawning may be induced in the laboratory through
manipulation of photoperiod and temperature allowing for a year-round supply of
reproductive mummichog (MacLatchy et al., 2003; MacLatchy et al., 2005). The egg has
a clear chorion allowing assessment of embryonic development (Armstrong and Child,
1965; Selman and Wallace, 1986). Mummichog have been reared successfully in the
laboratory and can reach sexual maturity in one year (40-50 mm) (Chapter 3; Matta et al.,
2001; Boudreau et al., 2004). Mummichog have been successfully used for EDS studies
in laboratory (MacLatchy et al., 2003; Sharpe et al., 2004), artificial stream (Dubé et al.,
2002), and field (Leblanc et al., 1997; Couillard and Nellis, 1999) assessments.
Page 50
Estuarine systems are essential spawning and nursery grounds for fish and other
vertebrates (Oberdortster and Cheek, 2000) and are frequently exposed to putative EDSs
from municipal and industrial discharges. Exposure in estuaries at sensitive life stages
such as gonadal recrudescence, embryonic and larval development could lead to
population-level impacts (Oberdortster and Cheek, 2000). Previous studies have shown
mummichog to be affected in the wild by the presence of xenobiotics such as pulp and
paper mill effluent (LeBlanc et al., 1997), mercury and petroleum hydrocarbons (Zhou et
al., 2000) and organics (Black et al., 1998). Laboratory studies have shown that different
life-cycle stages of the mummichog are sensitive to hormonally-active substances,
including adults (MacLatchy et al., 2003; Sharpe et al., 2004) and to a lesser extent,
embryos/larvae (Matta et al., 2001; Urushitani et al., 2002; Boudreau et al., 2004).
Society’s increased use of pharmaceutical estrogens has increased the presence of
estrogens in marine and estuarine environments (Oberdortster and Cheek, 2000). 17a-
ethynylestradiol (EE2) is a pharmaceutical (birth control pill and hormone replacement
therapy) that is not broken down in sewage treatment (Temes et al., 1999) and is a
reference chemical for EDS testing (OECD, 1999). Concentrations of EE2 present in
Canadian sewage treatment plants (STP) typically fall in the range of l-10ng/L EE2,
although levels have been documented as high as 42ng/L (Desbrow et al., 1998; Temes et
al., 1999). Estrogenic effects, including the presence of ovarian tissue in the testes, have
been demonstrated in fish downstream of STP outfalls (Harries et al., 1996; Jobling et al.,
1998; Larsson et al., 1999). In laboratory studies, EE2 has been chosen as a model EDS
for developing bioassays in mummichog due to its environmental relevance and because
Page 51
many EDSs exert their influence via estrogen receptor-mediated pathways (Ankley et al.,
2001; MacLatchy et al., 2003).
Recent short-term (7- or 15-day) exposures with mummichog revealed estrogenic
response and endocrine impacts at low, environmentally-relevant concentrations, as well
as similar responses at higher pharmaceutical concentrations of EE2 (MacLatchy et al.,
2003). However, no study has focused on the effects of EE2 exposure on the
reproductive potential of mummichog. Chronic exposure to xenobiotics can induce
changes to the organism not noted by shorter-term exposures, including factors that may
be manifested at the population level, such as lower recruitment (Ankley et al., 2001).
The objective of this present study was to develop a short-term reproductive
bioassay to link reproductive endocrine endpoints in adult fish to reproductive potential.
Circulating steroids (to indicate general endocrine effects), steroid production (to
determine biosynthetic capacity of gonadal tissue), vitellogenin (VTG) induction (to
measure estrogenic response), and fecundity and fertility (to measure reproductive
potential) were assessed after EE2 exposure of reproductively mature mummichog during
gonadal recrudescence and spawning (21- and 28-day exposures). This work furthers our
ability to extrapolate between short-term endocrine response and population-level
responses.
Page 52
Materials and Methods
Chemicals
17a-Ethynylestradiol (EE2; 98% purity) was purchased from Sigma-Aldrich
Canada (Oakville, ON, Canada). EE2 was stored at -20°C in 100% ethanol (Les Alcools
de Commerce, Boucherville, PQ, Canada) at stock concentrations of 3ng/mL, 30ng/mL,
300ng/mL and 3000ng/mL EE2 . Unless otherwise indicated, chemicals and reagents
were purchased from Sigma-Aldrich and laboratory supplies from Fisher Scientific
(Nepean, ON, Canada).
Fish
Adult mummichog were collected by seining from an uncontaminated estuarine
site (LeBlanc et al., 1997; Couillard and Nellis, 1999), Horton’s Creek, in the Miramichi
estuary (New Brunswick, Canada) in fall 2002 and acclimated to laboratory conditions at
the University of New Brunswick (Saint John, NB, Canada). The fish were fed
commercial crushed trout pellets (Corey Feed Mills, Fredericton, NB, Canada) until
satiation every day and supplemented every other day with Nutrafin™ floating pellets for
cichlids (Rolf C. Hagen, Montreal, QC, Canada). Fish were maintained in 250-L static,
filtered aquaria with dissolved oxygen (DO) at >85%, 15%o salinity (filtered Bay of
Fundy sea water and dechlorinated City of Saint John water), ambient temperature (18-
20°C) and natural photoperiod. Water quality testing (pH, temperature, ammonia, DO)
was completed regularly, and partial water changes were done periodically to maintain
standardized conditions. Cycle ™ (Rolf C. Hagen, Montreal, QC, Canada) was added
Page 53
when filters were cleaned and when water changes occurred to maintain beneficial
bacteria levels. Mortalities were minimal in the stock tanks (<5%).
Exposures
The range of EE2 concentrations were chosen to encompass environmentally-
relevant concentrations (synthetic and natural estrogens found in STP effluents)
(Desbrow et al., 1998). Concentrated EE2 was delivered in lmL of ethanol (0.0033%
EtOH/tank) to create final exposure concentrations of 0, 0.1, 1, 10, or lOOng/L EE2 .
In May 2003, spawning coloration was apparent in male fish (yellow bellies) and
female fish had begun to swell in the abdominal region. At this time, fish were removed
from the stock aquaria, separated by sex to minimize spawning events prior to egg
collection, and randomly allocated to 37-L static aquaria containing 30 L of 20%o saline
water (18-22°C; 16:8 h light:dark, DO > 80%) filtered by standard glass wool and
charcoal filters. Salinity in the experimental tanks was increased by 5%o as previous
spawning and/or larval studies with mummichog had been conducted at salinities of
greater than 18%o (Taylor et al., 1979; Hsiao and Meier, 1989; Matta et al., 2001;
Urushitani et al., 2002; Boudreau et al., 2004). Within each aquarium, fish were placed
inside a five-sided cage (no top) made from 3.175-mm plastic mesh screening (Aquatic
Ecosystems, Apopka, FL, USA) that was fitted to the sides of the tank and was suspended
2.5 cm above the bottom. Each aquarium contained twelve fish of a single sex and there
were ten tanks per sex. Fish were acclimated to these experimental conditions for two
weeks (8 May - 22 May 2003). The feeding regime was 3% body weight daily for the
Page 54
first week of acclimation and was reduced to 1.5% daily for the remainder of the
experiment as uneaten food was fouling the water.
The chemical exposure began on 22 May 2003, one week following the full
moon. At this point in the mummichog reproductive cycle, gonad size (GSI; [gonad
wt/body wt]-100) is typically low as the fish begin to develop mature gametes for the next
spawning period (Hsiao and Meier, 1989). On Day 1, the filters were removed and daily
renewal of water and treatment began. The fish were maintained in single-sex tanks for
the initial 21-day exposure period, with two tanks of each sex allocated to each treatment
group. Every morning, each basket was lifted and removed, with fish, and placed into a
spare aquarium containing 6 L of aerated 20%o water. The water in the treatment tank
was drained and replaced with 30 L of clean 20%o water. The basket containing the fish
was replaced into the original aquarium and EE2 treatment solution was added with a
pipette at the appropriate concentration. The ethanol solution was added near the airstone
to allow circulation of the EE2 prior to replacing the fish. The water was changed in
increasing treatment order (from Ong/L to lOOng/L) to avoid contamination. On Day 21
(12 June 2003), half of the fish were randomly selected from each tank for sampling of
reproductive endocrine endpoints. Mortality was 1.25% (three fish; one fish in each of
three tanks) in the first 21 days of EE2 exposure and there were no mortalities in the final
week of exposure.
Fecundity measures
On Day 21, a one-week reproduction trial began with the remaining fish. This
period was chosen to coincide with the full moon (Day 23 of the chemical exposure),
Page 55
which has been associated with peak GSI and spawning activities in mummichog
(Cochran, 1987; Hsiao and Meier, 1989). On Day 21, each male tank was matched with
a female tank of the same treatment and half of the males were exchanged with half of
the females, resulting in 20 tanks (four per treatment) each containing three males and
three females. Mummichog would naturally spawn against the mesh basket each day and
eggs would pass through the mesh to the bottom of the aquarium where they could not be
consumed by adult fish.
Each day, after the cage had been removed for water and treatment renewal,
naturally spawned eggs were collected by gently dredging the bottom of the aquarium
with a dip net (Urushitani et al., 2002). After all eggs were collected the water was
changed as described above. Egg collection occurred between 13 June and 19 June 2003.
After collection, the eggs were placed in clean, 20%o water and were examined at
4X magnification under a dissecting microscope. The total number of eggs spawned per
treatment, the mean spawning rates, the number of spawns per tank, and the number of
eggs per spawn were determined. A spawning event was defined as 10 or more eggs
found in a tank in one day. An estimation of fertilization was made based upon the
presence or absence of cell cleavage (Armstrong and Child, 1965), and nonviable or
unfertilized eggs were counted and removed. Fertilized eggs were placed in petri dishes
at a starting density of 30 eggs per dish and held in 50 mL of 20%o water in the same
treatment as their parents. Development of the embryonic, larval and juvenile stages was
further monitored through sexual maturation (Chapter 3).
Page 56
Fish sampling and reproductive endocrine endpoints
After 21 days of exposure, six fish were removed from each tank, except the three
tanks where there had been mortality, where only five fish were removed. These fish
were anesthetized (buffered 0.05% tricaine methane sulfonate; Syndel Laboratories,
Vancouver, BC, Canada), and measurements of weight and standard length were taken.
Blood samples were collected from the caudal vasculature with heparinized needles and
syringes (26 3/8 gauge needles on 1-mL syringes). Blood was mixed with aprotinin
(1KIU/|jL) upon collection and then centrifuged (2,400 x g at 4°C) for 12 min to isolate
plasma (Denslow et al., 1999; MacLatchy et al., 2005). Aprotinin is a protease inhibitor
that will slow the degradation of vitellogenin (Denslow et al., 1999). Plasma was divided
into two volumes; 10 jiL was stored at -80°C until vitellogenin (VTG) could be measured
by enzyme-linked immunosorbent assay (ELISA) and the remainder was stored at -20°C
for future steroid measurement by radioimmunoassay (RIA).
After killing the fish by spinal severance, the gonads and livers were dissected
and weighed to determine GSI and liver somatic index (LSI; [liver wt/body wt]-100).
Condition factor, a measure of body weight to standard length, was also calculated (CF;
[body wt/standard length3]-100). Gonadal tissue was placed in chilled Medium 199
(M l99) prior to preparation for in vitro incubation. An optimized protocol for measuring
mummichog gonadal steroid production in vitro was utilized (MacLatchy et al., 2005).
The testes were minced and two pieces (19-25g/mL total mass) were incubated (6 wells
per fish total) in polystyrene culture plates. Ovarian tissue was examined and pre-
maturational follicles (0.8-1.25 mm diameter [Selman and Wallace, 1986]; 19-25mg/mL
total mass) were incubated as per male tissue. In vitro incubations were 24 h at 18°C in
Page 57
fresh M l 99 (to synchronize incubation) supplemented with 1 mM of phosphodiesterase
inhibitor 3-isobutyl 1-methylxanthine (IBMX) (basal treatment; three wells), or with 1
mM IBMX and 20IU/mL human chorionic gonadotropin (hCG-stimulated treatment;
three wells). IBMX stimulates steroidogenesis by preventing the breakdown of cAMP
(3’,5’-cyclic adenosine monophosphate) and enhances steroid production in mummichog
incubations (McMaster et al., 1995; MacLatchy et al., 2003). Our laboratory uses steroid
production in the M l99 + IBMX samples as our “basal” production levels as amplified
steroid production can increase the chances of detecting in vivo treatment differences and
determining mechanisms of action of xenobiotic interference with the steroidogenic
pathway, though we recognize that this is not a ‘true’ basal level of steroid production
(MacLatchy et al., 2003). At 24 h, the incubation medium was collected with a pipette
without disturbing the gonadal tissue. The solution was frozen at -20°C until RIA was
used to determine levels of hormone production.
Radioimmunoassay fo r steroids
RIA was used to measure the concentrations of testosterone (T), 17|3-estradiol
(E2) and 11-ketotestosterone (11-KT) in the plasma and the in vitro medium using a
standardized protocol modified to be more precise for small plasma volumes typically
collected from the mummichog (20-50 |iL) (MacLatchy et al., 2005). Inter-assay
variability for T was 13.0%, E2 was 11.9% and 11-KT was 11.8%. Intra-assay
variability for each steroid was approximately 5%. Antibodies for T and E2 were
purchased from Medicorp (Montreal, QC, Canada), radiolabelled T and E2 from
Amersham Pharmacia Biotech (Baie D’Urfe, QC, Canada) and unlabelled T and E2 from
Page 58
Sigma-Aldrich. 11-KT antibody was from Helix Biotech (Vancouver, BC, Canada),
radiolabelled 11-KT from Dr. Timothy Gross (University of Florida, Gainesville, FL,
USA) and unlabelled 11-KT from Steraloids (Wilton, NJ, USA).
Vitellogenin assay
VTG levels in the plasma were determined using an ELISA for mummichog
(Denslow et al., 1999; MacLatchy et al., 2005). Mummichog VTG standard was
produced by Dr. Nancy Denslow (University of Florida, Gainesville, FL, USA). Primary
antiserum was produced by Dr. Charles Rice (Clemson University, Pendleton, SC, USA).
Secondary antibody was a peroxidase conjugate - goat anti-mouse IgG (Sigma-Aldrich).
Inter-assay variability was 14.5% and intra-assay variability was 6%.
Statistics
Statistical analyses were conducted using SigmaStat 3.0 (SPSS, Chicago, IL,
USA) or Systat 9.0 (Systat Software Inc., Richmond, CA, USA). Significant differences
(p<0.05) among the data were assessed separately for each sex. Hormonal data were
examined for outliers using Dixon’s test and only samples that had 30-40 |xL (female) or
15-20 |xL (male) plasma available for extraction were used. Prior to parametric analysis,
assumptions of normality and variance homogeneity were tested using normal probability
plots and Levene’s test. Where the data failed to meet the assumptions, they were logio
transformed and the assumptions retested. If the data still failed to meet the assumptions,
an equivalent non-parametric test was used. Analyses of weight, plasma and in vitro
steroids, and plasma VTG were performed using a nested analysis of variance (ANOVA)
Page 59
followed by Tukey’s or Dunn’s tests, with tank as the factor nested within treatment
group. Kruskal-Wallis was used when the data failed to meet assumptions. Differences
in gonadal and liver weights relative to body weights (GSI and LSI) and body weight to
standard length (CF) were determined using analysis of covariance (ANCOVA). The
interaction term significance was set at p=0.1 (Environment Canada, 1997). Where the
assumptions were not met for ANCOVA, regression lines were examined independently.
Measures of spawning and fertilization were assessed using single factor and two-factor
ANOVA. Values are reported as means ± standard error (SE).
Results
21-day exposure
There were no significant effects of EE2 on the weight of male (approximately 7 g;
p=0.084; Table 2.1) or female (approximately 12 g; p=0.52; Table 2.1) fish. Similarly
there were no differences in standard length of male (approximately 7.5 cm; p=0.188;
Table 2.1) or female (approximately 8.45 cm; p=0.99; Table 2.1) fish. There were no
differences in LSI in either sex (male p=0.43; female p=0.24; Table 2.1). Male GSI was
elevated in the lOOng/L EE2 group (4.23% of body weight) compared to the O.lng/L
group (3.33%; p=0.018; Table 2.1). CF was significantly lower in males exposed to
lng/L EE2 (p=0.013; Table 2.1). ANCOVA for female GSI and CF showed significant
interactions (p<0.1; data not shown). Under separate regression analysis, all treatments
showed increased gonad weight with body weight for female fish. Upon examination of
the data, it was apparent that one female in each treatment except lOOng/L was
leveraging the trends. When these individuals (four fish in total) were removed from the
Page 60
Table 2.1: Mean [±1SE] weight, length, liver somatic index (LSI), gonadosomatic index (GSI) and condition factor for adult male and
female Fundulus heteroclitus exposed to EE2 for 21 and 28 days during gonadal recrudescence (21 days) and spawning (28 days) in
June 2003. Means with different letters are significantly different (p<0.05).
Sex Exposure Length (d) Treatment (ng/L) N Weight (g) Length (cm) LSI (%) GSI (%) Condition
0 12 7.71±0.47a 7.52±0.14a 3.50±0.36a 3.50±0.21ab 1.80±0.03a
0.1 12 7.53±0.34a 7.63±0.08a 2.96±0.21a 3.33±0.18a 1.68±0.04abMale 21 1 11 6.90±0.46a 7.47±0.19a 2.79±0.20a 3.68±0.20ab 1.64±0.04b
10 12 6.31±0.31a 7.18±0.12a 3.00±0.16a 3.67±0.13ab 1.70±0.04ab
100 12 7.37±0.28a 7.46±0.11a 3.60±0.39a 4.23±0.16b 1.77±0.03ab
0 12 10.95±0.52a 8.44±0.16a 4.44±0.35a 12.19±1.46a 1.86±0.12ab
0.1 12 11.81±1.08a 8.43±0.27a 4.75±0.39a 13.33±1.19a 1.94±0.08abFemale 21 1 12 14.51±3.01a 8.44±0.44a 3.93±0.39a 16.25±1.59a 2.18±0.09b
10 10 11.12±0.97a 8.39±0.16a 4.07±0.24a 15.27±2.65a 1.86±0.09ab
100 12 11.59±0.96a 8.56±0.30a 4.42±0.26a 10.70±1.00a 1.83±0.07a
0 12 7.69±0.60a 7.58±0.20a 2.55±0.21a 2.61±0.15a 1.73±0.04aA 1U .l i
1Z•-1 , r \ '•ysa./ .JZ = tU .JZ 7.45±0.11a 2.60±0.19a 2.55±0.14a 1.76±0.04a
Male 28 1 12 7.71±0.34a 7.61±0.11a 2.44±0.18a 2.72±0.12ab 1.74±0.03a
10 12 6.94±0.34a 7.28±0.16a 2.42±0.13a 2.99±0.12ab 1.80±0.05a
100 12 7.58±0.44a 7.63±0.17a 2.34±0.18a 3.32±0.21b 1.69±0.04a
0 12 14.79±1.48a 8.98±0.96a 3.50±0.17a 12.15±1.73a 2.01±0.11a
0.1 12 13.00±1.03a 8.71±0.19a 3.74±0.25a 15.96±2.65a 1.95±0.09a28 1 12 12.28±1.16a 8.57±0.30a 4.28±0.20a 12.49±1.47a 1.86±0.05a
10 12 12.20±0.86a 8.61±0.23a 3.77±0.24a 11.54±2.57a 1.90±0.06a
100 12 11.41±1.44a 8.38±0.20a 3.74±0.29a 13.68±2.10a 1.87±0.08a
Page 61
data set, the ANCOVA assumptions could be met and GSI was not different among the
treatments (p=0.148; Table 2.1). Regression analysis of CF showed increased body
weight with body length in all treatments except Ong/L, which showed no change in body
weight with changes in length (WT = 1.171 - 0.148X; p==0.859, R2 = 0.000). Upon
examination of the data, it was apparent that one female was leveraging this trend. When
removed from the data set, the ANCOVA assumptions could be met and CF was
significantly higher at lng/L than lOOng/L (p=0.045, Table 2.1).
Exposure at lOOng/L induced a significant elevation in plasma VTG in male fish
(p<0.001; Fig 2.1 A). Female fish exposed to EE2 did not show any significant difference
in plasma VTG levels (p=0.55; Fig 2.1 A).
Exposure to EE2 did not affect circulating T levels in male mummichog at any
concentration (p=0.32) but circulating 11-KT levels were significantly depressed at 1, 10
and lOOng/L (p<0.001; Fig 2.2A). In female fish, plasma T and E2 levels were
significantly reduced at 100 ng/L EE2 (p=0.027; p=0.003; Fig 2.2B).
Following exposure to EE2, males showed reduced T production at lOOng/L (basal:
p<0.001; hCG: p=0.039; Fig 2.3A). There was a significant reduction in hCG-stimulated
T and E2 production in female fish exposed to 1 and 1 Ong/L EE2 (T: p=0.025; E2 :
p=0.014; Figs 2.3B and 2.3C) although basal treatments did not demonstrate these effects
(T: p=0.072; E2: p=0.107; Figs 2.3B and 2.3C).
28-day exposure
There were no significant effects of EE2 on the weight of male (approximately 7 g;
p=0.75; Table 2.1) or female (approximately 12 g; p=0.29; Table 2.1) fish. Similarly
Page 62
EE2 Treatment (/L)
Figure 2.1. Plasma vitellogenin (VTG) levels in male and female mummichog exposed to
0, 0.1,1,10 or lOOng/L 17a-ethynylestradiol (EE2) for 21 (A) or 28 (B) days. Bars
represent means ± SE. Means showing different letters are significantly different
(p<0.05). n=9-12 fish per group.
Page 63
0 0.1 10 100 0 0.1
EE2 Treatment (ng/L)
10 100
Figure 2.2. (A&C) Plasma testosterone (T) and 11-ketotestosterone (11KT) from male
mummichog exposed to 0, 0.1, 1, 10 or lOOng/L 17a-ethynylestradiol (EE2) during
gonadal recrudesence (2Id - A) or during spawning (28d - C). Bars represent means ±
SE, n=6-12 fish per group. Means showing different letters are significantly different
(p<0.05).
Page 64
IT3•oo55
'Si)&.2*CXA
EE2 Treatment (ng/L)
Figure 2.2. (B&D) Plasma T and 17(3-estradiol (E2) from female mummichog exposed to
0, 0.1, 1, 10, lOOng/L EE2 during gonadal recrudesence (2Id - B) or spawning (28d - D).
Bars represent means ± SE, n=6-12 fish per group. Means showing different letters are
significantly different (p<0.05).
Page 67
there were no differences in length of male (approximately 7.5 cm; p=0.51; Table 2.1) or
female (approximately 8.6 cm; p=0.56; Table 2.1) fish. There were no differences in LSI
(male p=0.84; female p=0.27; Table 2.1) or CF (male p=0.76; female p=0.66; Table 2.1)
in either sex. Male GSI was elevated in the 100ng/L EE?, group (3.32% of body weight)
compared to the 0.1ng/L (2.55%) and control (2.61%; p=0.006; Table 2.1) groups. There
were no differences in female GSI among treatments (p=0.52; Table 2.1) after 28 days.
VTG induction in males was significantly higher after exposure to 100ng/L EE2 for
28 days (p<0.001; Fig 2. IB). There were no differences in plasma VTG levels in female
fish exposed to EE2 for 28 days (p=0.058; Fig 2.IB).
There were no differences in circulating T or 11-KT levels in male fish after 28 days
of exposure (p=0.29; p=0.31; Fig 2.2C). T levels were lower at 28 days than at 21 days
of exposure (p<0.043; Figs 2.2A and 2.2C). In female fish there was no difference in
plasma T after 28 days of exposure (p=0.098; Fig 2.2D). E2 levels were significantly
depressed at 10 and 100ng/L (pO.OOl; Fig 2.2D). Overall female plasma hormone levels
were similar between 21 days and 28 days of exposure (Figs 2.2B and 2.2D).
After exposure to EE2 for 28 days, males exposed to 100ng/L EE2 showed
significantly lower T production in basal testes incubations (p=0.003; Fig 2.4A). In
hCG-stimulated incubations, T production was lower compared to the lng/L treatment
only (p=0.001; Fig 2.4A). Basal incubations of ovarian tissue showed a significant
increase in T production in fish exposed to 1 and 10ng/L EE2 and in E2 production in fish
exposed to lng/L EE2 for 28 days (T: p=0.002; E2: p=0.024; Figs 2.4B and 2.4C). In
hCG-stimulated incubations, there was also a significant increase in T production by
female fish exposed to 1, 10 and 100ng/L EE % compared to control (p=0.003; Fig 2.4B).
Page 70
There was no change in E2 production among the groups except for between the 1 and
100ng/L treatments (p=0.016; Fig 2.4C).
Fecundity Measurements
At the end of the egg collection period (Day 22-28), females in the 100ng/L tanks had
spawned significantly fewer eggs than those in Ong/L and other EE2 treatment tanks
(p=0.037; Fig 2.5). The 100ng/L fish had a lower mean spawning rate over the 7-day
collection (p=0.031) than all other treatment groups [35.7±6.1 (Ong/L), 43.1±8.5
(0.1ng/L), 39.3±9.8 (lng/L), 44.4±5.0 (10ng/L) and 15.6±4.3 (100ng/L) eggs per female
per day]. The number of eggs per spawning event was not different among the treatment
groups (p=0.184, data not shown) and the difference in fecundity may be linked to the
number of spawning events - 6.5±0.3 (Ong/L), 5.8±0.3 (0. lng/L), 4.9±0.9 (lng/L),
6.0±0.7 (1 Ong/L) and 3.5±1.0 (100ng/L) - where the 100ng/L tanks had fewer spawning
events than the 0, 0.1 and 1 Ong/L tanks (p=0.048). Fertilization success, as measured by
percentage of eggs fertilized, was significantly lower at 100ng/L than Ong/L (p<0.001;
Fig 2.6).
Page 71
Cum
ulati
ve
num
ber
of eg
gs s
pawn
ed
per
fem
ale
21 22 23 24 25 26 27 28 29
Time (days from full moon)Exposure Day
Figure 2.5. Mean cumulative eggs spawned per female inummichog exposed to 0, 0.1,1,
10 or 100ng/L 17a-ethynylestradiol. Means showing different letters are significantly
different (p<0.05). n=4 tanks per treatment.
Page 72
T3<u
cdPh</3
3«4HoP5O
'£oo,2&e_oOSN
tÌ<D
EE, Treatment (ng/L)
Figure 2.6. Mean fertilization rate of mummichog eggs calculated daily between days 21
and 28 of a 28-day exposure to 0, 0.1,1,10 or 100ng/L 17a-ethynylestradiol. Bars
represent means ± SE. Means showing different letters are significantly different
(p<0.05). n=4 tanks per treatment.
Page 73
Discussion
In this study, mummichog exposed to EE2 at greater than environmentally-
relevant concentrations (lOOng/L) exhibited disruption to normal endocrine function and
reproduction. However, exposure to lower concentrations of EE2 (<10ng/L) resulted in
minimal disruption in the reproductive endpoints measured. This work grew out of
previous studies in our labs (MacLatchy et al., 2003; Boudreau et al., 2004) examining
effects in adult and larval mummichog exposed to EE2 . The goal of the present study
was to determine if a standardized multigenerational protocol for EDS studies in
mummichog could be used to correlate physiological effects with population and/or
reproductive impacts as has been suggested for other fish species (Patyna et al., 1999;
Ankley et al., 2001). The long-term goal is to be able to tie biochemical and/or whole
organism effects to population effects in order to allow for the shortening of the full life
cycle test; however, further work is required to fully understand the links between
endocrine responses and reproductive effects.
Vitellogenin (VTG) induction in male and/or juvenile fish has been used as a
biomarker indicative of estrogenic exposure (Denslow et al., 1999) in freshwater (Patyna
et al., 1999; Ankley et al., 2001) and saltwater (Karels et al., 2003; MacLatchy et al.,
2003) fish. Induction of VTG has been demonstrated in short-term (Panter et al., 2002),
long-term and/or multigenerational studies (Ankley et al., 2001; Karels et al., 2003; Seki
et al., 2003). In the present study, plasma VTG was elevated in males exposed to
lOOng/L EE2 for 21 and 28 days (Table 2.2). In a recent short-term exposure study, male
mummichog exposed to lOng/L and lOOng/L for 15 d and 50ng/L for 7 d had increased
Page 74
Table 2.2: Summary of changes in plasma vitellogenin, plasma titres and gonadal production o f testosterone (T), 11-ketotestosterone
(11-KT; males only) and 17P-estradiol (E2; females only) and spawning (females only) and fertilization (males only) in mummichog
exposed to graded doses of waterborne ethynylestradiol (EE2) for 21 or 28 days. <-» = no significant response; | = increase; | =
decrease; * = variable not measured. Numbers beside symbols represent the concentrations (ng/L) at which the changes occurred.
**Fertilization success may also be attributable to ova or an ova-sperm combination.
Exposure Length (d) Sex VTG
PlasmaT T Production
Plasma 11- KT Plasma E2
E2Production
CumulativeSpawning Fertilization**
21 6 100 f 100 i 1 , 10,1004 * * * *
$ «—>• 1 0 0 | 1 , 1 0 4 * 100 4 l, 1 0 4 * *
28 6 l o o t <r-> 100 4 * * * 100 i
?1 , 10, 100 t * io, ioo 4 1 , l o t 100 i *
■o
Page 75
plasma VTG (MacLatchy et al., 2003). Those fish were 2-3 g heavier than those used in
the present study and it is possible that the larger livers associated with the larger fish had
more estrogen receptors resulting in an increased vitellogenin response. In a preliminary
multi-generational study in our lab, male mummichog exposed to environmentally-
relevant concentrations of EE2 for 56 days had increased VTG only at 100ng/L
(unpublished data). Environmentally relevant concentrations of EE2 may not be enough
to elicit a consistent estrogenic response, or regulatory mechanisms may be more
effective over a longer exposure time. Effects on VTG have been found at concentrations
as low as 4ng/L in juvenile fathead minnow; however, sex was not determined in these
fish (Länge et al., 2001). In a short-term sheepshead minnow study, VTG induction was
only observed at 100ng/L in males (Folmar et al., 2000). Within species, effects of EE2
exposure may depend upon life stage, reproductive stage, or length of exposure and there
appear to be species-specific sensitivities.
In female mummichog, serum and ovarian E2 levels, and oocyte maturation,
parallel the semilunar spawning cycle (Taylor and Dimichelle, 1980; Bradford and
Taylor, 1987). The gene for VTG is activated in mature females once a threshold level of
E2 is reached in the hepatic estrogen receptors (Denslow et al., 1999). VTG is produced
in the liver and circulated in the blood until it is sequestered into the ovary. It is
transformed into yolk proteins during the major growth phase of oocyte development
(Selman and Wallace, 1986; Denslow et al., 1999; Kime et al., 1999). In this study,
females exposed to EE2 for 21 and 28 days did not show any differences in plasma VTG
(Table 2.2). The sampling dates bracketed the peak spawning time, when gonadal E2
production and circulating E2 levels are at a maximum (Bradford and Taylor, 1987). In a
Page 76
recent study investigating the effects of EE2 exposure, VTG was found to increase in a
concentration-dependant manner after 7- and 15-day exposures (MacLatchy et al., 2003).
The data were collected at different points in the lunar cycle from the present study.
When E2 production and serum levels are at naturally low levels, it is possible that
exogenous estrogen can more effectively compete for hepatic estrogen receptors (Kime et
al., 1999). This binding could maintain the threshold level required for VTG production
resulting in detectable differences in circulating VTG. Exogenous estrogen effects may
diminish during peak oocyte developmental periods when ovarian E2 production
increases and VTG production and sequestration are already at maximum inducible
levels. Lack of VTG induction may also be the result of feedback and regulatory
physiological mechanisms within the fish achieving homeostasis over a longer exposure
period (Länge et al., 2001).
In recrudescing fish (21-day sampling event), male fish exposed to EE2 showed
depressions in gonadal T production (100ng/L) and plasma 11-KT (l-100ng/L) (Table
2.2). Depressions in gonadal steroid production were not manifested in plasma T levels,
indicating that homeostatic mechanisms were regulating circulating levels or that
exposure duration was not long enough to result in changes in circulating T due to
reduced gonadal production. Alterations in levels of circulating sex steroid levels are a
whole organism response indicating potential effects on steroid availability at target cells.
Therefore, circulating levels can be considered a more significant indicator of potential
long-term effects on reproductive status than gonadal production. Depressions in plasma
11-KT levels could have been due to lack of T substrate in the gonad; however, 11-KT
production in the testes incubations was not measured. Female fish exposed to 100ng/L
Page 77
EE2 exhibited depressions in both plasma E2 and T on Day 21; both T and E2 production
was also decreased at intermediate concentrations (1 and lOng/L) (Table 2.2).
Results from the Day 28 fish demonstrate effects of EE2 exposure during the
active spawning period (Table 2.2). In males, overall circulating levels of T were lower
after 28 days of exposure compared to the levels on Day 21 and may have been due to
drops in gonad size during spawning. T production remained depressed at lOOng/L.
Differences in plasma 11-KT observed at Day 21 were diminished on Day 28.
Reproductive cycle variation in plasma 11-KT levels in mummichog is well documented
(Cochran, 1987); effects of EE2 during recrudescence and spawning may reflect a
difference in sensitivity of 11-KT production and/or control during these different
periods. The results were consistent with findings in MacLatchy et al. (2003) in which
males exhibited depressions in circulating androgens and gonadal steroid production after
short-term exposure to EE2, and the magnitude and/or presence of changes changed with
the lunar cycle. In females, reductions in plasma T in lOOng/L fish at Day 21 were not
apparent on Day 28 (Table 2.2). The most dramatic difference between Day 21 and Day
28, however, was in gonadal steroid production. T production was increased by EE2
treatments >1 ng/L and E2 production was increased at 1 and lOng/L EE2 (Table 2.2). In
a previous study (MacLatchy et al., 2003), steroid production was decreased at high
exposure concentrations (250-1000ng/L) in a 7-day study and E2 production was
increased at exposure concentrations of 1 to lOOng/L EE2 after 15 days of exposure.
Changes in steroid production patterns are probably the r esult of different sensitivities
during various periods of the female reproductive cycle which can respond in either a
positive or negative feedback fashion. As well, female fish have homeostatic
Page 78
mechanisms in the endocrine system for both estrogens and androgens which enable a
tight balance of plasma steroids (Balch et al., 2004).
Further study is needed to determine whether steroid production and plasma
levels are generally parallel or inconsistent in the mummichog. Interpretation of the in
vitro incubation results are difficult in this study because there was no T, 11-KT, or E2
response to hCG stimulation in male or female fish. Previous studies with mummichog
in our laboratory have demonstrated successful stimulation of gonadal steroid production
at 20 IU/mL hCG (MacLatchy et al., 2003; Sharpe et al., 2004). HCG stimulation
amplifies the response and also indicates whether there may be a problem with
gonadotropin-mediated stimulation of steroidogenesis. The effectiveness of this
particular batch of hCG is being investigated as control fish were not stimulated. In
addition, the impact of the presence or absence of the opposite sex on steroid levels and
initiation of spawning requires further investigation, as this may have contributed to
differences independent of programmed reproductive cycling. Relative sensitivity issues
at different parts of the reproductive cycle need to be further addressed through detailed
studies during different reproductive stages in mummichog.
Mummichog eggs were collected over a one-week period (Day 22 - Day 28)
beginning the day before the full moon, as preliminary studies indicated that the majority
of fish had completed spawning 2-7 days after the full moon (unpublished data).
Fecundity and % fertilization were chosen as our key indicators of reproductive success.
The addition of these endpoints to our short-term endocrine bioassay (MacLatchy et al.,
2003; Sharpe et al., 2004) addresses the concern that altered physiological and/or
biochemical status may not manifest directly as impaired reproductive success (Ankley et
Page 79
al., 2001). Female fish exposed to 100ng/L EE2 spawned significantly fewer eggs than
all other treatments. This trend has been found in other species exposed to EE2 and other
estrogens. Fathead minnows had reduced egg production after exposure to methoxychlor
and E2 (Patyna et al., 1999; Ankley et al., 2001); Japanese medaka had fewer copulations
and reduced egg production at 10ng/L EE2 (Balch et al., 2004) and decreased egg
production at 500ng/L EE2 (Tilton et al., 2005); a reduction in viable eggs was found in
sheepshead minnow exposed to 4-tert-octylphenol as concentration increased (Karels et
al., 2003); and EE2 decreased reproductive success in sheepshead minnow at 200 ng/L
(Zillioux et al., 2001). Mummichog, therefore, are consistent with other fish species in
which estrogenic compounds decrease egg production in short-term and life-cycle
bioassays.
Once the extreme outliers were removed, no differences were observed in female
GSI among the treatments on each day, or between Day 21 and Day 28 (Table 2.1). Prior
to a spawning peak, female GSI is usually 12-14% and falls to 1-7% post-spawning
(Taylor et al., 1979; Taylor and DiMichele, 1980). After the egg collection period, all
treatments had mean GSIs above 10% and upon examination of the ovaries subsequent to
dissection many contained mature oocytes indicating that not all fish had spawned. Five
of the Ong/L fish and five of the 0.1 ng/L fish at Day 28 had GSIs greater than 15%
(compared to two fish at 1 and 1 Ong/L and three fish at 100ng/L), which indicates the
potential for continued spawning in control and 0.1 ng/L treatments had the study not
been terminated. This difference in timing of spawning between the treatments and the
alteration of gonadal steroid production at 1 and 1 Ong/L on both Day 21 and Day 28
indicate that low concentrations of EE2 inay induce gonadal development and/or affect
Page 80
reproductive cycling; this requires further investigation in mummichog. Stimulation of
reproductive capabilities at low concentrations of EE2 and inhibition at high
concentrations has been observed in Japanese medaka (Tilton et al., 2005).
Declines in fertility have been noted in sheepshead minnow at 200ng/L (Zillioux
et al., 2001) and Japanese medaka at lOng/L (Balch et al., 2004) EE2. Fertilization
success may be affected by either gamete, and it is possible that egg quality affected
fertilization in the present study. However, effects on fertilization in fish have
predominately been linked with effects on males. EE2 exposure with medaka has shown
increased testis cell death and no impact on ovarian cell death (Weber et al., 2004) and no
change in ovarian architecture or oogenesis (Balch et al., 2004). Reproductive studies
with male sheepshead minnow exposed to 4-teri-octylphenol showed that fertilization
success decreased with increased concentration when they were mated with control
females, indicating that fertilization success is largely dependent on the male gamete
(Karels et al., 2003). Male rainbow trout had reduced fertilization success at 10 and
lOOng/L EE2 accompanied by increased sperm concentration that was probably the result
of inhibited seminal fluid production (Schultz et al., 2003). Whether reduced fertilization
in mummichog is due to effects on males requires further study; the observed reduction in
fertilization indicates that there may be disruption to gonadal development or
spermatogenesis in the testes in the lOOng/L EE2 exposure as GSI was increased in males
in this treatment (Table 2.1). However, no major abnormalities were noted in casual
observation of the testes during sampling. Increased GSI indicates accelerated gonadal
growth, a lack of sperm release or an inhibited maturation cycle. Feminization of male
fish through inhibited or abnormal testes growth (including development of ovarian
Page 81
tissue) (Jobling et al., 1998; Gill et al., 2002) and altered secondary sex characteristics
(Länge et al., 2001; Parrott and Wood, 2002) is a common effect of estrogen exposure,
though the concentration at which feminization occurs depends on species, length of
exposure, the estrogenic chemical and life stage at which exposure began. Feminization
or absence of the vas deferens has been seen in carp (Cyprinus carpio) exposed to 17ß-
estradiol in the laboratory and in roach (Rutilus rutilus) exposed to STP effluents in the
wild (Gimeno et al., 1998; Jobling et al., 1998). A feminized or absent vas deferens
could block the release of viable gametes during spawning (Gill et al., 2002), which
would cause reduced fertilization and a temporarily increased GSI until feedback
mechanisms inhibited further development and/or initiated atresia. It is doubtful that our
short-term exposure of normally recrudescing adult males that have already initiated
spawning would have resulted in structural changes; effects on maturation and/or release
of sperm are more likely.
In summary, low levels of EE2 (0.1-10ng/L), such as those found in the Canadian
environment, do not appear to impair reproductive function in mummichog after short
term exposure, despite some disruption of biochemical and physiological endpoints (this
study; MacLatchy et al., 2003). VTG induction, the endocrine system and biosynthetic
pathways, fecundity and fertilization were affected in mummichog exposed to 100ng/L
EE2 . These results indicate that there may be a link between altered endocrine, status and
reproductive problems; further investigation may complete the link and provide
mechanistic explanations. Fathead minnow and Japanese medaka have shown strong
impacts of EE2 at environmentally-relevant concentrations (Länge et al., 2001; Parrott
and Wood, 2002; Balch et al., 2004) when long-term studies commenced exposure prior
Page 82
to reproductive maturity. Though current EDS studies with mummichog have assessed
critical endpoints in individual life stages (Matta et al., 2001; MacLatchy et al., 2003;
Boudreau et al., 2004), only completion of the life-cycle bioassay for mummichog will
link short-term and longer-term EDS studies with common physiological, biochemical,
reproductive and developmental endpoints. The eggs resulting from this study were
exposed to EE2 through hatching, and offspring were reared under the same treatments to
complete the life cycle, providing further insight into the effects of chronic estrogen
exposure on developing mummichog (Chapter 3). This reproductive bioassay is also
being used to assess the potential of industrial discharges to affect reproductive potential
in this species.
Acknowledgements
This study was funded by a Discovery Grant to DLM from the Natural Sciences and
Engineering Research Council of Canada as well as a Networks of Centres of Excellence:
Canadian Water Network grant to DLM (PI: K. Munkittrick, UNB Saint John). At UNB
Saint John, the following technicians and students are thanked for their assistance with
the exposure bioassays and laboratory work: J. Adams, M. Beyea, C. Blanar, K.
Gormley, J. Ings, L. Peters, R. Sharpe, K. Shaughnessy, G. Vallieres, and L.Vallis. Drs.
N. Denslow (University of Florida) and C. Rice (Clemson University) are thanked for
their assistance in developing the mummichog vitellogenin ELISA.
Page 83
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Chapter 3*: Early-life Development and Sex Differentiation
Effects of 17a-ethynylestradiol on early-life
development, sex differentiation and vitellogenin
induction in mummichog (Fundulus heteroclitus)
*This chapter is to be submitted to the journal Ecotoxicology and Environmental Safety
under joint authorship with Simon C. Courtenay (Fisheries and Oceans Canada) and
Deborah L. MacLatchy (University of New Brunswick).
Page 93
Abstract
Fertilized mummichog eggs retrieved from 17a-ethynylestradiol (EE2) exposed
adult fish were raised in concentrations of EE2 ranging from 0 to lOOng/L for 61 weeks
post-hatch. Eggs exposed at lOOng/L hatched sooner and larvae were longer than all
other treatments, though fewer hatched at this treatment. Survival of juvenile fish from
hatch to termination of the study was higher at 1 OOng/L EE2 treatment. There were no
differences found in growth or vertebral abnormalities. Sex ratios were found to be
skewed toward females at lOOng/L EE2, and some gonadal male fish displayed female
secondary sex characteristics. Condition factor, gonadosomatic index (GSI), and liver
somatic index (LSI) were found to decrease between 48 and 61 weeks post-hatch. Males
had decreased GSI at lOOng/L at 52 weeks post-hatch. Female fish had increased hepatic
vitellogenin (VTG) and there were no differences in male hepatic VTG at 52 weeks post
hatch. Males exposed at lOOng/L may have had disruption at some endpoints (GSI, VTG)
that is masked due to reduced sample size compared to other treatments. Fish exposed to
concentrations of EE2 below environmentally relevant levels (i.e., <10ng/L) showed
inconsistent effects on development and reproductive potential. Individual reproductive
indices with the potential to produce changes at the population level were evident after
chronic exposure of developing mummichog at lOOng/L EE2.
Key Words
Ethynylestradiol, Fundulus heteroclitus, vitellogenin, sex differentiation, development,
lifecycle, bioassay, endocrine disruption
Page 94
Introduction
Impacts of endocrine disrupting substances (EDSs) on fish (e.g., on characteristics
such as survival, morphology, tissue organization, and fecundity) and their offspring (eg.,
on hatching, growth, survival, and sex ratios) have been demonstrated in lifecycle
bioassays (Patyna et al., 1999; Zillioux et al., 2001; Parrott and Wood, 2002). Chronic
exposure to xenobiotics can induce changes to organisms not noted by shorter-term
exposures, including factors that may be manifested at the population level, such as lower
recruitment (Ankley et al., 2001). Lifecycle bioassays or multigenerational bioassays
have been developed in freshwater fish species, including the fathead minnow,
Pimephalespromelas (Ankley et al., 2001; Länge et al., 2001), Japanese medaka, Oryzias
laripes (Seki et al., 2003) and zebrafish, Danio rerio (Olsson et al., 1999). Partial and
short-term bioassays have been developed for estuarine and marine species such as
sheepshead minnow, Cyprinodon variegates (Folmar et al., 2000; Zillioux et al., 2001)
and mummichog, Fundulus heteroclitus (MacLatchy et al., 2003; Boudreau et al., 2004).
Mummichog are numerically dominant in salt marshes along the east coast of
Canada and the United States (Armstrong and Child, 1965), and have demonstrated
sensitivity to endocrine disrupting substances EDSs in laboratory (MacLatchy et al.,
2003; MacLatchy et al., 2005), artificial stream (Dube et al., 2002), and field (Leblanc et
al., 1997) assessments. The mummichog is a good candidate for lifecycle bioassay
development due to its ease of husbandry and reproductive biology. Mummichog are
relatively sedentary, exhibiting small home ranges in the wild (Lotrich, 1975) and they
are potentially exposed to environmental EDSs throughout their life cycle.
Page 95
In laboratory studies, 17a-ethynylestradiol (EE2) has been chosen as a model EDS
for developing bioassays in mummichog due to its environmental relevance, as well as its
confirmed effects on the reproductive endocrine system via estrogen receptor-mediated
pathways (OECD, 1999; Ankley et al., 2001; Metcalfe et al., 2001; MacLatchy et al.,
2003). EE2 is one component of sewage effluent associated with increased incidence of
intersex in male fish exposed downstream of sewage treatment plants (STPs) (Jobling et
al., 1998). It is a synthetic pharmaceutical (birth control pill and hormone replacement
therapy) that is not broken down in sewage treatment; concentrations of EE2 present in
Canadian STPs are usually between l-10ng/L EE2, although levels have been
documented as high as 42ng/L (Desbrow et al., 1998; Temes et al., 1999) and is diluted
downstream by river water and rain water (Harries et al., 1996). In earlier short-term (7-
or 15-day) exposure studies using EE2, adult mummichog displayed endocrine impacts at
low, environmentally-relevant concentrations, as well as similar responses at higher
pharmaceutical concentrations (MacLatchy et al., 2003). In longer-term (21- or 28-day)
EE2 exposures, reproductive cycling was shifted in females, sex steroid production and
circulating levels were altered and at lOOng/L EE2, fecundity and fertility were reduced
(Chapter 2).
The objective of this study was to determine the impact of chronic EE2 exposure
on offspring development. Embryos derived from parents exposed during pre-spawning
and spawning phases were continuously exposed to EE2 for 15 months, through their
development to pre-spawning juveniles/adults. Embryonic/larval endpoints (time to
hatch, hatch success, length at hatch), larval/juvenile endpoints (growth, survival,
vertebral abnormalities) and yearling endpoints (liver vitellogenin, gonad and
Page 96
liversomatic indices, condition factor, and sex ratios) were evaluated for developmental
(embryonic, larval and juvenile stages) and reproductive (yearlings) anomalies. This
study, in conjunction with our previous studies (MacLatchy et al., 2003, Boudreau et al.,
2004; MacLatchy et al., 2005; Boudreau et al. in press) furthers our ability to understand
the effects of EDSs on various lifestages of mummichog.
Materials and Methods
Chemicals
The 17a-ethynylestradiol (EE2; 98% purity) was purchased from Sigma-Aldrich
Canada (Oakville, ON, Canada). EE2 was stored at -20°C in 100% ethanol (Les Alcools
de Commerce, Boucherville, PQ, Canada) at stock concentrations of 3ng/mL, 30ng/mL,
300ng/mL and 3000ng/mL EE2 for adult exposures and lOng/mL, lOOng/mL, lOOOng/mL
and lOOOOng/mL EE2 for larval and juvenile exposures. Unless otherwise indicated,
chemicals and reagents were purchased from Sigma-Aldrich and laboratory supplies from
Fisher Scientific (Nepean, ON, Canada).
Exposures
Adult fish were collected by beach seine from an uncontaminated estuarine site
(LeBlanc et al., 1997; Couillard and Nellis, 1999), Horton’s Creek, in the Miramichi
estuary (New Brunswick, Canada) in fall 2002 and acclimated to laboratory conditions at
the University of New Brunswick (Saint John, NB, Canada). Concentrated EE2 was
delivered in ethanol (0.0033% EtOH/tank for adults; 0.001% EtOH/tank for offspring) to
create final nominal exposure concentrations of 0, 0.1,1,10, or lOOng/L EE2 .
Page 97
Environmentally-relevant concentrations of EE2 (O-lOng/L) were chosen based upon
synthetic and natural estrogen levels found in STP effluents (Desbrow et al., 1998).
Details of the adult exposure protocol have been reported previously (MacLatchy et al.,
2003; MacLatchy et al., 2005). Fertilized eggs were collected between 13 June and 19
June 2003 (Day 22-Day 28 of adult exposure) from experimental tanks fitted with a five
sided cage (no top) made from 3.175-mm plastic mesh screening (Aquatic Ecosystems,
Apopka, FL, USA). Each aquarium contained six fish (three males; three females) and
there were four tanks per treatment (twenty tanks in total). Adult mummichog placed
inside the cage would naturally spawn against it, allowing eggs to pass through the mesh
to the bottom of the aquarium where they could not be consumed. Each day, after the
cage (and adult fish) had been removed for water and treatment renewal, naturally
spawned eggs were collected by gently dredging the bottom of the aquarium with a dip
net (Urushitani et al., 2002).
Exposures - offspring
After collection, the eggs were placed in clean, 20%o saline water (filtered Bay of
Fundy sea water and dechlorinated City of Saint John water) and examined for
fertilization at 4X magnification under a dissecting microscope. Fertilization was
determined by the presence or absence of cell cleavage (Armstrong and Child, 1965).
Fertilized eggs were retained and transferred to glass petri dishes at a starting density of
30 eggs per dish and held in 50 mL of 20%o salinity in the same treatment as their parents
(0, 0.1, 1, 10, or lOOng/L EE2).
Page 98
The dishes were held unoxygenated at ambient temperature (18-22°C) and
photoperiod (16:8 L:D) on a laboratory bench until the eggs hatched. Each dish was
examined daily at 4X magnification, and dead embryos and hatched larvae were
removed. After daily examination, the water was removed from each dish and replaced
with 50 mL of water treated with the appropriate amount of EE2. Time to hatch, survival
to hatch and length at hatch were monitored over the hatching period.
Upon hatch, larvae were transferred to 20%o saline water in 50mL beakers
containing the same EE2 concentrations at a maximum density of 10 larvae per beaker.
Larvae were held at ambient temperature (18-22°C) and photoperiod (16:8 L:D) with
daily water changes until the yolk sac was absorbed (1-3 days) and swimming began.
The larvae were then randomly allocated to one of four aerated, static 37-L aquaria per
treatment (twenty tanks in total). Every morning, all fish were removed from each tank
with a dip net and placed into a spare aquarium containing 6 L of aerated 20%o water.
The water in the treatment tanks was drained and replaced with the appropriate amount of
clean 20%o saline water. The fish were then replaced into the original aquarium and EE2
treatment solution was added with a pipette at the appropriate concentration. The ethanol
solution was added near the airstone to promote circulation of the EE2. The water was
changed in increasing treatment order (from Ong/L to lOOng/L) to avoid contamination.
Upon initiation of the growout phase, each aquarium contained 5 L of EE2-treated 20%o
saline water (18-20°C, 16:8 h L:D, dissolved oxygen>80% saturation). At five weeks,
the volume of water in each tank was increased to 10 L to accommodate growing larvae.
Beginning at 10 weeks, the volume of water was adjusted separately for each tank to
minimize the effect of density differences among the tanks due to differential survival.
Page 99
Water volume was set to 1 L per 1 g total wet weight of the fish in the tank and was
adjusted every 3-4 weeks for the remainder of the experiment. At week 23, to
accommodate growing fish, the 37-L aquaria were replaced with 75-L aquaria. Due to a
dosing error in week 14, replicate tanks for Ong/L and lOOng/L EE2 were reduced to three
tanks for subsequent endpoints.
Larvae were fed live newly hatched Artemia sp. nauplii (Bohai Bay Salt Ponds
Artemia Cysts, Aquatic Ecosystems, Apopka, FL, USA) enriched with Roti-rich™
(Aquatic Ecosystems) twice daily (1 mL of concentrated Artemia per L of water) and fry
food (Rolf C. Hagen, Montreal, QC, Canada) to satiation once daily for 14 weeks.
Beginning at 8 weeks, freeze-dried Red Grubs (Rolf C. Hagen) were used to supplement
the Fry Food diet as the juveniles were weaned from the Artemia. Flaked Staple Food
(Rolf C. Hagen) was introduced as the primary feed at 22 weeks, fed 2-3 times daily,
supplemented by Red Grubs or Cichlid Food (Rolf C. Hagen) once daily.
Photoperiod was adjusted to simulate seasonal day lengths, thus providing the fish
with visual cues for gonadal development. The photoperiod was set at 16:8 h light:dark
from the initiation of the experiment until week 14 (July-October), when it was reduced
to 14:10 h light:dark until week 20 (October-November). At week 20 it was reduced to
12:12; light:dark and remained at this period until week 32 (November-February). At
week 32, the photoperiod was increased to 14:10 light:dark until week 40 (February-
April) when it was increased to 15:9 light:dark (April-May). At week 46 the photoperiod
was increased to 16:8 light:dark, where it remained until the conclusion of the
experiment. Temperature was held at ambient conditions, which decreased slightly from
Page 100
18-21°C in the summer (weeks 1-20) to 16-18°C for the winter (weeks 21-40), returning
to 18-21°C for the remainder of the study.
Growth was assessed weekly for the initial seven-week period, and then
approximately every three weeks for the remainder of the experiment. Growth was
determined by collecting total length measurements (to the nearest mm) from 25 fish
scooped from each tank. Survival was determined from the number of fish in each tank
on each measurement day, and calculated as the proportion of the original number of fish
in the tank. Vertebral abnormalities were assessed at weeks 15, 48 and 61 and ranged
from mild (one or two slight bends in the spine) to severe (one or more bends in the spine
that dramatically altered body shape and/or affected swimming ability) and included
scoliosis (lateral curvature) and lordosis (dorsoventral curvature) (Boudreau et al., 2004).
At week 48 (May 2004), all fish were assessed for weight (balance to 0.00lg),
standard and total length (mm), vertebral abnormalities and external sex. Sex was
determined based upon colouration and secondary sex characteristics including yellow
bellies, dark dorsal fin spot, silver vertical stripes, white or yellow spots in males, and
brown or green backs and pale bellies in females (Armstrong and Child, 1965). At week
52 (June 2004), 20 fish were randomly sampled from each tank except one lng/L and one
lOOng/L where 0 and 15 fish were selected respectively, due to low numbers of fish in
these tanks. These fish were anaesthetized (buffered 0.05% tricaine methane sulfonate;
Syndel Laboratories, Vancouver, BC, Canada), assessed for weight (to 0.00lg), standard
and total length (mm), vertebral abnormalities and sex based upon secondary
characteristics and gonadal assessment at 4X magnification.
Page 101
Vitellogenin (VTG) in the liver was also assessed at week 52 using an ELISA
modified from MacLatchy et al (2004). Liver tissue was dissected out, weighed, frozen
on dry ice, and then stored at -80°C. For each mg of liver tissue, 1 |xL of aprotinin
(1KIU/ |aL) was added to the microfuge tube. Tissue was thawed and homogenized with
a Kontes Pellet Pestle hand-held homogenizer. A small amount of the slurry (5|xL) was
diluted 100X with Tris buffered saline containing Tween and bovine serum albumin and
assayed as per MacLatchy et al (2004). Prior to analysis, the method was tested by
spiking liver slurry samples with a known concentration of VTG. Recovery of the spiked
amount of VTG [recovered VTG = 0.549 + 1.071 * (spiked VTG); R2=0.774; p<0.001]
was 88%. Inter-assay variability was 7% and intra-assay variability was 6%.
At week 61 (August 2004), all remaining fish were sampled and assessed for
weight, standard and total length, vertebral abnormalities and sex based upon secondary
characteristics and gonadal assessment at 4X magnification.
Statistics
Statistical analyses were conducted using SigmaStat 3.0 (SPSS, Chicago, IL,
USA) or Systat 9.0 (Systat Software Inc., Richmond, CA, USA). Significant differences
(p<0.05) among the data were assessed separately for each sex when it was possible to
determine gonadal sex. Prior to parametric analysis, assumptions of normality and
variance homogeneity were tested on morphological data using normal probability plots
and Levene’s test. Where the data failed to meet the assumptions, they were logio
transformed and the assumptions retested. If the data still failed to meet the assumptions,
an equivalent non-parametric test was used. Measures of time to hatch, hatch success
Page 102
rate, and length at hatch were assessed using single factor ANOVA, using Petri dishes as
the replication units. Predicted growth rates were determined from the von Bertalanffy
growth curves (Cailliet et al., 1986). Mortality, vertebral abnormalities and liver
vitellogenin (VTG) were assessed using single factor ANOVA, with tank as the unit of
replication. Vitellogenin data were examined for outliers using Dixon’s test prior to
analysis. Analyses of weight and length were performed using a nested analysis of
variance (ANOVA) followed by Tukey’s or Dunn’s tests, with tank as the factor nested
within treatment group. Kruskal-Wallis was used when the data failed to meet parametric
assumptions. Differences in gonadal and liver weights relative to body weights
(gonadosomatic index; GSI; [gonad wt/body wt]-100; and liversomatic index; LSI; [liver
wt/body wt]-100), body weight to standard length (condition factor; CF; [body
wt/standard length3]-100), and total length over time (growth) were determined using
analysis of covariance (ANCOVA). The interaction term significance was set at p=0.1
(Environment Canada, 1997). Differences in morphological variables between sampling
dates were assessed using a student’s t-test. Sex ratios were assessed using a chi-square
analysis for proportions, with expected proportions set at 50:50, followed by an angular
transformation and Tukey-type multiple comparison (Zar, 1999). Values are reported as
means ± standard error (SE).
Results
Eggs exposed to lOOng/L EE2 hatched sooner than eggs exposed to 0, 0.1, or
lng/L EE2, and eggs exposed to lOng/L EE2 hatched sooner than those exposed to Ong/L
(p=0.018; Fig. 3.1). Eggs exposed to lOOng/L EE2 had significantly lower hatch success
Page 104
Total Length (mm)
-D 3Hatch Success (% of fertilized eggs) Time to Hatch (days)
oa
3era
3era
aera
o3era
oo3era
00-J
Page 105
than those in 0, 0.1, and lng/L (p<0.001; Fig. 3.1). Larvae hatched at 1 and 100ng/L EE2
were significantly longer than those at 0, 0.1 and 10ng/L (p<0.001; Fig. 3.1).
The predicted growth rate from the von Bertalanffy growth curves showed a
similar trend among treatments (Fig. 3.2A) and total length did not differ among
treatments from hatch to 61-weeks post-hatch (p=0.340; Fig. 3.2B). Weight at length
(condition factor) for male fish was not different among treatments at week 48, 52 or 61
(Table 3.1) and decreased between week 48 and week 61 in all treatment groups (Table
3.2). In females, weight at length was not different between Ong/L and any other
treatment at week 48, 52, or 61, though differences occurred among some treatments at
week 48 and week 61 (Table 3.1). Female condition decreased between week 48 and
week 61 in all treatment groups (Table 3.2).
Weight and length of male fish exposed to 1 Ong/L EE2 were significantly lower
than other treatments of lesser concentrations at 48 weeks post-hatch (Table 3.1). At
100ng/L, only length was significantly lower (Table 3.1). These effects had disappeared
at 52 and 61 weeks post-hatch (Table 3.1). At 48 weeks post-hatch, female fish exposed
to 1 Ong/L EE2 were found to be shorter and weigh less than those exposed to all other
treatments (Table 3.1). Length and weight effects had diminished at 52 and 61 weeks
(Table 3.1), with the exception of female fish exposed to 100ng/L EE2 who were heavier
than fish exposed to 0.lng/L EE2 at 61 weeks post-hatch (Table 3.1).
Post-hatch mortality was found to be significantly lower for 100ng/L exposed fish
than any other treatment, and fish exposed to lng/L had higher mortality than fish
exposed to 0.1 and 1 Ong/L (pO.OOl; Fig. 3.3). At week 15 the proportions of fish
exhibiting vertebral abnormalities at 0.1,10, and 100ng/L were significantly lower than
Page 107
Total
Len
gth
(mm
) W
Proj
ected
To
tal
Leng
th
(mm
)
Time (Weeks Post-Hatch)
Page 109
Sex ExposureLength
Treatment(ng/L) n Weight (g) Length (cm) LSI (%) GSI (%) Condition
0 82 0.84±0.03a 3.52±0.05a - - 1.93±0.05a
0.1 96 0.79±0.03a 3.49±0.05a - - 1.84±0.05a
Male 48 weeks 1 80 0.81±0.03a 3.48±0.06a - - 1.97±0.09a
10 119 0.67±0.02b 3.27±0.04b - - 1.90±0.06a
100 5 0.59±0.08ab 3.02±0.25b - - 2.26±0.38ap-value 0.001 0.001 - - 0.215
0 104 0.85±0.06a 3.50±0.06a - - 1.95±0.05ab
0.1 131 0.78±0.03a 3.48±0.05a - - 1.83±0.05a
Female 48 weeks 1 136 0.84±0.03a 3.44±0.05a - - 2.11±0.07b
10 180 0.66±0.03b 3.17±0.04b - - 2.00±0.04ab
100 176 0.87±0.03a 3.51±0.04a - - 1.95±0.05abp-value 0.001 0.001 - - 0.014
0 30 0.70±0.05a 3.44±0.09a 2.59±0.18a 1.42±0.15a 1.67±0.07a
0.1 41 0.75±0.05a 3.44±0.07a 2.84±0.22a 1.63±0.39a 1.70±0.06a
Male 52 weeks 1 26 0.68±0.06a 3.22±0.12a 2.97±0.36a 1.29±0.13a 1.96±0.12a
10 38 0.75±0.04a 3.40±0.06a 3.05±0.28a 1.36±0.17a 1.87±0.05a
100 7 0.75±0.14a 3.41±0.19a 3.83±0.46a 0.60±0.23b 1.74±0.08ap-value 0.715 0.293 0.486 0.016 0.089
0 29 0.88±0.07a 3.71±0.11a 3.36±0.23a 4.30±0.87a 1.66±0.06a
0.1 38 0.82±0.07a 3.57±0.10a 3.78±0.30a 2.55±0.14a 1.73±0.07a
Female 52 weeks 1 34 0.91±0.06a 3.65±0.08a 3.57±0.24a 3.10±0.32a 1.83±0.06a
10 42 0.84±0.06a 3.54±0.08a 3.33±0.21a 3.11±0.46a 1.76±0.06a
100 47 0.96±0.05a 3.76±0.07a 3.86±0.22a 4.81±0.99a 1.76±0.06ap-value 0.431 0.354 0.648 0.641 0.265
0 26 0.85±0.06a 3.76±0.08a 2.35±0.16a 0 .86±0 .12a 1.57±0.07a
0.1 37 0.71±0.05a 3.55±0.08a 1.97±0.12a 0.77±0.08a 1.50±0.05a
Male 61 weeks 1 51 0.85±0.05a 3.70±0.07a 1.98±0.13a 0.78±0.07a 1.60±0.04a
10 34 0.83±0.06a 3.72±0.07a 2.33±0.17a 0.82±0.10a 1.54±0.06a
100 12 0.65±0.07a 3.53±0.15a 2.27±0.35a 0.45±0.13a 1.46±0.10ap-value 0.155 0.497 0.242 0.395 0.418
0 27 0.94±0.05ab 3.91±0.08a 2.45±0.19ab 2.78±0.77a 1.56±0.05ab
0.1 42 0.85±0.05a 3.82±0.08a 2.59±0.15ac 1.78±0.09a 1.49±0.07a
Female 61 weeks 1 45 0.98±0.05ab 3.93±0.08a 2.05±0.14b 2.07±0.19a 1.58±0.04ab
10 38 0.91±0.06ab 3.83±0.07a 2.85±0.10a 3.62±0.79a 1.57±0.04ab
100 47 l.ll±0 .06b 4.04±0.07a 2.24±0.11bc 3.56±0.81a 1.65±0.06bp-value 0.025 0.371 <0.001 0.364 0.024
Page 110
Table 3.2. P-values for changes in morphological variables of juvenile mummichog
among weeks 48, 52, and 61 post-hatch after exposure to 0, 0.1 ,1 ,10 or 100ng/L 17a-
ethynylestradiol (EE2). Arrows indicate the direction of change where significant.
ee2(ng/L)
48-52weeks
Male52-61weeks
48-61weeks
48-52weeks
Female52-61weeks
48-61weeks
p-value p-value p-value p-value p-value p-value0 0.38 0.018t 0.017t 0.074 0.147 o.ooit
0.1 0.618 0.303 0.442 0.414 0.035t <0.00 it
Standard Length 1 O.O274 <0.0011 0.02t 0.037t o.oit <0.0011
10 0.072 o.ooit o.ooit 0.002t 0.009t o.ooit
100 0.236 0.663 0.099 0.0031 0.007Î o.ooit
0 0.0084 0.027t 0.925 0.723 0.503 0.221
0.1 0.476 0.526 0.15 0.842 0.407 0.19
Weight 1 0.0434 0.052t 0.895 0.212 0.386 0.013Î10 0.067 0.319 0.0031 0.009t 0.748 o.ooit
100 0.406 0.514 0.617 0.067t 0.049Î o.ooit
0 O.OI24 0.069 <0.0014 O.OOI4 0.234 <0.0014
0.1 0.712 O.OOI4 <0.0014 0.431 <0.0014 <0.0014
Condition Factor 1 0.927 O.OO34 <0.0014 0.0464 <0.0014 <0.0014
10 0.735 <0.0014 <0.0014 0.332 <0.0014 <0.0014
100 0.147 0.073 0.0134 0.091 0.0154 <0.0014
0 - 0.0094 - - O.OO64 -0.1 - <0.0014 - - <0.0014 -
GSI 1 - O.OOI4 - - <0.0014 -10 - O.OO24 - - 0.268 -
100 - 0.546 - - 0.074 -0 - 0.337 - - 0.0074 -
0.1 - 0.0044 - - <0.0014 -LSI 1 - O.OI4 - - <0.0014 -
10 - 0.095 - - 0.113 -100 - 0.0154 - - <0.0014 -
Page 111
Surv
ival
(%
aliv
e)
Time (weeks)
Figure 3.3. Mean survival of mummichog exposed to 0, 0.1, 1,10, or 100ng/L 17a-
ethynylestradiol from hatch to 61 weeks post-hatch, n =: 3-4 tanks per treatment.
Page 112
at 0 and lng/L (p=0.021; Fig. 3.4). At 48 weeks post-hatch, fish exposed to O.lng/L
exhibited significantly more vertebral abnormalities than fish exposed to 10 and lOOng/L
(p=0.045; Fig. 3.4) and at 61 weeks post-hatch all differences had diminished among the
treatments (p=0.502; Fig. 3.4). The proportion of fish exhibiting vertebral abnormalities
increased over time in fish exposed to O.lng/L EE2 (p=0.001; Fig. 3.4) and decreased
over time in fish exposed to Ong/L EE2 (p=0.032; Fig. 3.4) and did not change in the
other treatments (p=0.805; p-0.192; p=0.548; Fig. 3.4). Within each sex there was no
difference in vertebral abnormalities by treatment at week 48 (males p=0.384; females
p=0.437; data not shown) and at week 61 (males p=0.145; females p=0.650; data not
shown).
Male fish exposed to lOOng/L EE2had significantly lower GSI at 52 weeks but
not at 61 weeks post-hatch (Table 3.1) and between week 52 and week 61, male GSI
decreased in all treatments except lOOng/L EE2 (Table 3.2). No differences were found
in male LSI among treatments at 52 weeks or 61 weeks post-hatch (Table 3.1) and LSI
decreased between week 52 and week 61 in fish exposed to 0.1, 1 and lOOng/L EE2
(Table 3.2). There was no difference in female GSI between treatments at 52 weeks or
61 weeks (Table 3.1) post-hatch though female GSI decreased between week 52 and
week 61 in fish exposed to 0, 0.1, and lng/L EE2 (Table 3.2). No differences were found
between treatments in female LSI at 52 weeks post-hatch and although female LSI at
Ong/L was not different from any treatments at 61 weeks post-hatch, differences were
found among treatments (Table 3.1). Between week 52 and week 61 post-hatch, female
LSI decreased in all treatments except 1 Ong/L (Table 3.2).
Page 113
0.5
C/3
<4 -1OeiooaoSh
a 0.4
0.3
<D
o£
J-H
•s
>
0.2 -
0.1
0.0
ab
be
aI ac
b
ibe
beI
0.1 1 10
EE2 Treatment (ng/L)
Week 15 Week 48 Week 61
-be
100
Figure 3.4. Mean prevalence of vertebral abnormalities in juvenile mummichog exposed
to 0, 0.1, 1,10, or lOOng/L 17a-ethynylestradiol at 15,48 and 61 weeks post-hatch. Bars
represent means ± SE. n=3-4 tanks per treatment. Means showing different letters are
significantly different (p<0.05).
Page 114
At 52 weeks post-hatch, there was no difference among treatments in liver VTG
in males (p=0.732; Fig. 3.5). There was an increase in liver VTG in females exposed to
lOOng/L EE2 compared to all other treatments (p<0.001; Fig. 3.5).
At 48, 52, and 61 weeks post-hatch, sex ratios based on secondary sexual
characteristics were significantly different at lOOng/L, with females accounting for
greater than 90% of the fish at this concentration (48: p<0.001; 52: p<0.001; 61: p<0.001;
Table 3.3). Visual examination of the gonads at 52 and 61 weeks post-hatch found sex
ratios to be significantly different at lOOng/L EE2, the proportion of females being 86.8
and 81.5%, respectively (p<0.001; pO.OOl; Table 3.3).
Discussion
Previous studies in our laboratories (MacLatchy et al., 2003; Boudreau et al.,
2004; Chapter 2) examined the effects of EE2 exposure on adult and larval mummichog.
This present work addresses questions that arose from these studies, including whether a
standardized, multigenerational protocol could be developed for EDS studies to correlate
physiological effects with population and/or reproductive impacts in mummichog as has
been suggested for other fish species (Patyna et al., 1999; Ankley et al., 2001). The eggs
used in this study were spawned and fertilized in a short-term reproductive bioassay in
which adult mummichog were exposed to concentrations of EE2 ranging from 0 to
lOOng/L for 21 and 28 days (Chapter 2). The rearing of these eggs to sexual maturity
provides a more complete understanding of the total impact of chronic EE2 exposure.
Long-term exposure to the highest concentration (lOOng/L) resulted in a higher
Page 115
fcH<u>
20000
a 15000 H 00
e
& looooo"q
5000 -
Male (p=0.732) Female (p<0.001)
L à^-fc ■ ■ IOng 0.1 ng 1 ng 10 ng
EE, Treatment (/L)
100 ng
Figure 3.5. Liver vitellogenin levels in male and female mummichog exposed to 0,0.1,
1, 10 or lOOng/L 17a-ethynylestradiol (EE2) for 52 weeks post-hatch. Bars represent
means ± SE. Means showing different letters are significantly different (p<0.05). n = 2-
4 tanks per treatment.
Page 116
proportion of fish with female secondary sex characteristics and gonads. As well,
100ng/L for 21 and 28 days (Chapter 2). The rearing of these eggs to sexual maturity
provides a more complete understanding of the total impact of chronic EE2 exposure.
Long-term exposure to the highest concentration (100ng/L) resulted in a higher
proportion of fish with female secondary sex characteristics and gonads. As well, female
livers contained higher VTG levels at this concentration. These EE2-induced changes
occurred only at the highest concentrations and parallel the physiological and
reproductive effects observed in the parents of these offspring (decreased circulating
steroids, increased VTG, decreased egg production and decreased fertility; Chapter 2).
In the progeny generation, hatching success was reduced when parents and eggs were
exposed to 100ng/L EE2. Decreases in both spawning output and fertilization success
during adult exposure to 100ng/L EE2 have been demonstrated in mummichog in our
laboratory (Chapter 2). Thus, exposure of mummichog to EE2 during breeding at
concentrations greater than those found in the environment has significant negative
implications for total number of offspring produced. This may be a threshold effect, as
there is no indication of similar difficulties in spawning, fertilization or hatching at EE2
levels <10ng/L. However, a concentration-dependent response may exist between
exposure concentrations of 10 and 100ng/L EE2; to determine this requires further study.
Sheepshead minnow exposed to EE2 showed reduced egg production at 20 and 200ng/L
and reduced hatching at 200ng/L (Zillioux et al., 2001). Hatching success decreased in a
dose-dependent manner in larval mummichog exposed to estrogen (E2) (Urushitani et al.,
2002) and estrogenic alkylphenols caused complete larval mortality at concentrations >10
Page 117
Table 3.3. Proportion (mean ± 1SE) of juvenile mummichog exhibiting female
secondary sex characteristics and/or female gonads at 48, 52, and 61 weeks post-hatch
after exposure to 0, 0.1, 1, 10 or 100ng/L EE2. Differing superscripts identify
significantly different treatments within sampling date (p<0.05). n=3-4 tanks per
treatment.
Treatment(ng/L)
Exposure length N
Mean % $ ± SE
Secondary Sex Characteristics Gonadal Sex
0 3 0.559 ±0.003a -
0.1 4 0.588 ± 0.039a -
1 48 weeks post-hatch 4 0.631 ±0.014a -
10 4 0.601 ± 0.030a -
100 3 0.976 ±0.016b -
0 3 0.514 ±0.104a 0.495 ± 0.095a
0.1 4 0.462 ± 0.078a 0.480 ± 0.045a
1 52 weeks post-hatch 3 0.594 ± 0.075a 0.567 ± 0.058a
10 4 0.580 ± 0.084a 0.525 ±0.103a
100 3 0.928 ± 0.043b 0.868 ± 0.076b
0 3 0.534 ±0.053a 0.515 ±0.055a
0.1 4 0.574 ± 0.039a 0.546 ± 0.029a
1 61 weeks post-hatch 4 0.491 ±0.039a 0.455 ± 0.034a
10 4 0.566 ±0.038a 0.523 ± 0.040a
100 3 0.970 ± 0.030b 0.815 ±0.050b
Page 118
|o,M and hatching success was <10% below at this concentration (Kelly and Di Giulio,
2000).
As EE2 concentrations increased, there was a decrease in the egg incubation
period prior to hatch. Exposure to estrogens and/or estrogen-like compounds has not
been shown to change time to hatch in Japanese medaka. (Seki et al., 2003; Balch et al.,
2004), though shortened time to hatch has been documented in zebrafish exposed to the
weakly estrogenic pesticide methoxychlor (Versonnen et al., 2004) and brown trout
(Salmo trutta) exposed to wood-sterols (Lehtinen, et al., 1999). Previous work with
mummichog exposed to EE2 found no treatment effects on time to hatch, and increases in
length at hatch only in fish exposed at lOng/L, not at higher concentrations (Boudreau et
al., 2004). In the present study, larvae that emerged were significantly longer when
exposed to 1 and lOOng/L EE2, Boudreau et al (2004) did not assess lng/L, but found no
differences at lOOng/L EE2 . The inconsistency in results between the two studies might
be due to the initiation of EE2 exposure. In the present study, exposure began in the
parent generation, while in Boudreau et al (2004) exposure began post fertilization.
Exposure to EE2 prior to spawning and fertilization may affect the developing oocyte
before the development of the protective chorion and associated membranes of the egg
that prevent exogenous compounds from reaching the developing embryo (Anadu et al.,
1999).
Growth patterns of mummichog larvae and juveniles based upon total length did
not differ among treatments. Differences in total length were seen in both sexes at
lOng/L at 48 weeks post-hatch, but had diminished by 52 and 61 weeks post-hatch.
Measures of growth based upon length have shown growth is severely reduced at lng/L
Page 119
EE2 in fathead minnows (Länge et al., 2001), but not in sheepshead minnow exposed to
EE2 (Zillioux et al., 2001) or zebrafish exposed to methoxychlor (Versonnen et al., 2004).
Condition factor has also been used in similar studies as an indicator of growth. In this
study, there were no differences in male condition at 48, 52 or 61 weeks post-hatch.
Female condition in treated tanks was never different from control on any sampling date,
yet differences existed among treatments at 48 and 61 weeks (Table 3.1). Condition was
not affected in juvenile zebrafish exposed to methoxychlor (Versonnen et al., 2004), or
juvenile sheepshead minnows exposed to EE2 (Zillioux et al., 2001). Mummichog
exposed to EE2 from fertilization to 60 days post-hatch and then raised in clean water had
increased condition compared to controls at 6 and 12 months. In this case, increased
condition was linked to sexual maturity (Boudreau et al., 2004). In the present study,
between week 48 and week 61, condition factor decreased in all treatments in both sexes.
GSI also decreased in most treatments. Condition over time in fish is linked to the
reproductive cycle (increases with recrudescence, decreases with gonadal regression)
(Kleinkauf et al., 2004). Condition factor is also associated with the nutritive state of fish
(Saborowski and Buchholz, 1996). Deficiencies in diet have been shown to affect
mummichog growth which may affect reproductive maturity (Gutjahr-Gobell, 1998);
more work is required to optimize the diet of mummichog held in captivity and we
cannot assess whether improved nutrition would alter growth and development.
Mortality from hatch to the conclusion of the study was found to be significantly
lower for 100ng/L exposed fish than any other treatment. Fathead minnows exposed to
EE2 from fertilization showed no differences in mortality among treatments (Länge et al.,
2001). Mummichog exposed for 60 days and reared in clean water showed decreased
Page 120
mortality at lOng/L, no change in mortality at lOOOng/L and increased mortality at
10,000ng/L EE2 (Boudreau et al., 2004). In the present study, the increased survival in
the lOOng/L treatment tanks may be explained by a different density in these tanks. For
the initial 10 weeks of the growout period, all the aquaria held equivalent volumes of
water, however, the initial number of fish in the lOOng/L tanks ranged from 137 to 159,
and the initial number in the rest of the treatments (O-lOng/L) ranged from 230 to 318
fish. Boudreau et al. (2004) suggested that decreasing density by increasing volume of
water and/or decreasing number of fish reduced mortalities in mummichog, an effect that
seems to have been repeated in the current study. It is also possible that the decreased
mortality post-hatch in the lOOng/L treatment is the result of increased mortality as
embryos prior to hatch.
Differences in vertebral abnormalities observed among treatments at week 15
were not present at week 48 or 61. The proportion of fish exhibiting abnormalities did
not change among weeks 15,48 and 61 in the highest three treatments. The development
of abnormalities may have been delayed at 0. lng/L, resulting in the increase in the
proportion of affected fish with time. The proportion of Ong/L fish exhibiting vertebral
abnormalities decreased over time, and may be attributable to the mortality of these fish.
The growth of soft-tissue that hides underlying abnormalities has also previously been
suggested as a mechanism which decreases observable deformities (Boudreau et al.,
2004).
Vitellogenin (VTG) production in the liver naturally occurs in female fish once a
threshold level of E2 is reached in the hepatic estrogen receptors (Denslow et al., 1999).
Exogenous estrogen will induce VTG synthesis in female and male and/or juvenile fish
Page 121
(Denslow et al., 1999) in both freshwater (Panter et al., 2002) and saltwater (MacLatchy
et al., 2003) species. There was no difference in liver VTG in male fish at week 52.
However, the mean VTG for males exposed to 100ng/L was more than twice the mean of
male fish exposed to Ong/L EE2 . There were only five fish identified as male in the
100ng/L tanks, compared to 23-39 fish in each of the other treatments. These five fish
were only in two of the three 100ng/L replicate tanks. The small sample size for male
fish at this treatment level may be masking any effect. Alternatively, during a long
exposure, biofeedback and regulatory physiological mechanisms may allow the fish to
achieve homeostasis and prevent a VTG response (Länge et al., 2001). Females exposed
to 100ng/L EE2 had increased liver vitellogenin compared to all other treatments.
Medaka exposed to two estrogenic alkylphenols for 60 days post-hatch had
concentration-dependent increases in hepatic VTG in both sexes (Seki et al., 2003).
Whole body VTG increased in fathead minnows exposed to EE2 for 142 days post-hatch
and histological assessment of these individuals showed ovarian tissue in each animal
(Länge et al., 2001). Adult female mummichog showed elevated plasma VTG at
100ng/L EE2 in a 15-day exposure (MacLatchy et al., 2003), indicating consistent effects
through different life stages in this species.
Male fish exhibited no treatment differences in LSI at either 52 or 61 weeks post
hatch (Table 3.1). At week 52, there was no treatment difference in the LSI of female
fish and at week 61 post-hatch, there were no differences in LSI between Ong/L and any
of the EE2 treatments, though differences between the treatments existed. Liver sizes in
fish generally increase during vitellogenesis (Kleinkauf et al., 2004); however, there was
no correlation between LSI and liver VTG in this study (data not shown).
Page 122
At 52 weeks post-hatch, males exposed to lOOng/L EE2 had significantly lower
GSI than all other treatments, yet at 61 weeks post-hatch, there was no difference in male
GSI among treatments. However, at week 61, males exposed to lOOng/L EE2 had GSIs
less than 60% of each of the other treatments. Male GSI decreased between week 52 and
week 61 in tanks exposed to 0, 0.1, 1 and 1 Ong/L EE2. Few males showed the brilliant
yellow of sexual maturity and it is possible that male GSI was decreasing due to the
annual reproductive cycle (week 61 was 26 August 2004). GSI in mummichog decreases
as the breeding season ends (Cochran, 1987). Male fish exposed to lOOng/L EE2 did not
demonstrate the reduction in GSI between weeks 52 and 61 seen in the other treatments.
This may be due to sample size, as the number of male fish at lOOng/L was much
reduced. Male zebrafish exposed for 24 days at 10 and 25ng/L EE2 showed reductions in
GSI, due to a predominance of immature sperm (Van den Belt et al., 2002). Whether
sperm development was affected in mummichog requires further study.
There was no difference in GSI among treatments in female fish at 52 or 61
weeks post-hatch. Sampling at 52 weeks occurred in June, less than a week after the new
moon and at 61 weeks sampling occurred in August, four days before the full moon.
Mummichog have the highest GSI levels prior to spawning at the full moon (Taylor and
DiMichele, 1980). However, female GSI decreased between week 52 and week 61 in
fish exposed to 0,0.1, and lng/L EE2, and did not change in female fish exposed to 10
and lOOng/L EE2 (Table 3.2). If these fish were following the annual reproductive cycle,
the decrease in GSI indicates that higher levels of EE2 may cause differences in gonadal
development during the spawning season. In the parental generation, there was also
indication that EE2 exposure could cause disruption to reproductive cycles in female
Page 123
mummichog. Adult females exposed to higher concentrations of EE2 (1-100ng/L) had
altered gonadal steroid production and completed the spawning cycle earlier than fish
exposed to 0 and 0. lng/L EE2 (Chapter 2).
Sex ratios were determined based upon secondary sexual characteristics at 48, 52,
and 61 weeks post-hatch. Greater than 92% of the fish exposed at 100ng/L were female
and at all other treatment levels, the ratio did not deviate significantly from the expected
50% female. Visual examination (4X magnification) of the gonads at 52 and 61 weeks
post-hatch confirmed a skewed sex ratio at 100ng/L, though some fish classed as females
based upon secondary characteristics were actually gonadal males and the proportion of
females was shifted to approximately 81% at 100ng/L based upon gonadal tissue. This
method does not allow for assessment of intersex in these fish, and it is possible that
some of the fish classed as female in this study are males that have not been feminized
into fully functional females. Future histological assessment of all juvenile gonadal
tissue will determine the degree of feminization in these fish; however this lies beyond
the scope of this thesis. Early life exposure of mummichog to EE2 weakly skewed the
sex ratio towards females, even when subsequently raised in clean water (Boudreau et al.,
2004). Medaka eggs injected with EE2 have shown sex reversal of genetic males upon
growout (Papoulias et al., 1999) and medaka exposed to EE2 from 2 days post-hatch had
gonadal intersex at 2ng/L (Balch et al., 2004). Fathead minnows have been completely
feminized in lifecycle exposures to 4ng/L EE2 (Länge et al., 2001) and at 3.2ng/L in
exposures from the egg stage (Parrott and Wood, 2002). Mummichog sex ratios appear
to be affected at higher exposure concentrations than other species. Gonadal sex change
in fish can be induced by exposure to estrogens during windows of sensitivity related to
Page 124
gonadal differentiation and development. The fathead minnow has a period of enhanced
sensitivity to EE2 between 10 and 15 days post-hatch where male fish are more likely to
develop female characteristics (Van Aerie et al., 2002) and 6-day exposures with medaka
have demonstrated that exposure while an embryo, newly hatched fry or 1 week post
hatch are sufficient to alter sex ratios and/or develop intersex gonads at sexual maturity
(Koger et al., 2000). Little is known in mummichog about windows of enhanced EDS
sensitivity in regard to gonadal differentiation and this requires further study.
EE2 is typically found in low concentrations (0.1-10ng/L) in Canadian STP
effluents (Desbrow et al., 1998; Temes et al., 1999). Completion of the lifecycle
bioassay for mummichog has demonstrated that at environmentally-relevant
concentrations, exposure to EE2 during different life stages results in minimal, if any,
adverse effects on reproductive endpoints (this study; MacLatchy et al., 2003; Boudreau
et al., 2004). However, chronic exposure at 100ng/L EE2 negatively affected larval and
juvenile endpoints of hatch success, liver VTG, male GSI, and sex ratios in mummichog.
Long-term studies with fathead minnow and Japanese medaka exposed to
environmentally-relevant concentrations of EE2 have shown strong impacts when
exposure began prior to reproductive maturity (Länge et al., 2001; Parrott and Wood,
2002; Balch et al., 2004).
More studies are required to determine the relative sensitivities of different fish
species, including those typically classified as “model” species. Fathead minnows have
demonstrated failure to develop secondary sex characteristics, altered mating behaviour,
and VTG induction in males at concentrations of 2-4ng/L EE2 (Länge et al., 2001; Lattier
et al., 2002); Japanese medaka have demonstrated intersex and reduced copulations at
Page 125
concentrations of 2-10ng/L EE2 (Balch et al., 2004); and zebrafish have demonstrated
reduced spawning, fertilization and GSI at 1 Ong/L EE2 (Van den Belt et al., 2001; Van
den Belt et al., 2002). These freshwater species have all demonstrated significant
reproductive impacts at concentrations <1 Ong/L EE2. Sheepshead minnow have
demonstrated reduced reproductive success at concentrations of EE2 between 20 and
200ng/L (Zillioux et al., 2001) while mummichog have demonstrated impacts on steroid
synthesis, VTG induction, reproductive success and sex differentiation at lOOng/L EE2
(MacLatchy et al., 2003; Chapter 2; this study). These estuarine species appear to have a
higher tolerance for EE2 exposure than freshwater species studied. Differences in effects
may be due to physiological differences between estuarine and freshwater species and/or
differences in availability or uptake of contaminants; this should be further investigated.
Of most concern in this study are endpoints with the potential to affect the
population structure. Chronic exposure of successive generations of mummichog to EE2
resulted in reduced spawning, fertilization, hatching and altered sex ratios. This implies
fewer fish being added to the population with each successive year, and coupled with
altered GSIs in adult and juvenile males, indicates that the problem could compound with
time. We are presently validating these results with industrial discharges and other model
EDSs to determine the potential long-term impacts of endocrine disruption we have
observed in short-term adult bioassays (MacLatchy et al., 2003; Sharpe et al., 2004).
Development of the reproductive bioassay and growout protocols described here will be
of benefit when studying the impacts of other contaminants.
Page 126
Acknowledgements
This study was funded by a Discovery Grant to DLM from the Natural Sciences and
Engineering Research Council of Canada as well as a Networks of Centres of Excellence:
Canadian Water Network grant to DLM (PI: K. Munkittrick, UNB Saint John). At UNB
Saint John, the following technicians and students are thanked for their assistance with
the exposure bioassays and laboratory work: J. Adams, M. Beyea, C. Blanar, K.
Gormley, J. Ings, L. Peters, R. Sharpe, K. Shaughnessy, G. Yallieres, and L.Yallis.
Page 127
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MacLatchy, D.L., Gormley, K.L., Ibey, R.E.M., Sharpe, R.L., Shaughnessy, K.S.,
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Chapter 4: General Discussion
The effects of chronic exposure to 17a-ethynylestradiol on the
lifecycle of mummichog (.Fundulus heteroclitus)
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This study combined a mummichog short-term adult reproductive bioassay, early
life-stage bioassay and growout protocol. The adult reproductive bioassay assessed
endocrine and reproductive endpoints in adult mummichog exposed to 0, 0.1, 1, 10 or
100ng/L EE2 for 21 and 28 days. The early life-stage bioassay tracked the offspring from
the adult bioassay, exposed to the same concentrations as their parents, and assessed
embryonic survival and hatching. Growout continued the exposure to 61 weeks post
hatch and assessed growth and reproductive potential of the offspring.
The short-term adult reproductive bioassay assessed endpoints currently used in
short-term laboratory EDS studies and added measures of spawning and fertilization as
indicators of reproductive potential. This study integrated short-term biochemical and/or
whole organism effects that reflect EDS exposure with reproductive effects that may
carry forward to future generations (Chapter 2). Minimal disruption to the reproductive
endocrine system occurred in fish exposed at environmentally relevant concentrations of
EE2 (<10ng/L) in both the 21-day and 28-day exposures. However, there were
indications from alterations in female steroid production, GSI and spawning, that low
level EE2 exposure may induce gonadal development and/or affect reproductive cycling
in female mummichog. Exposure to the highest concentration of EE2 tested (100ng/L)
induced VTG production in male adults and resulted in consistent endocrine changes in
both sexes. Female fish exhibited decreased circulating steroid levels (testosterone and
estradiol) and increased testosterone production. Male fish exhibited decreased
testosterone production and decreased circulating 11-ketotestosterone. These results
were consistent with previous short-term exposures of adult mummichog to EE2
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(MacLatchy et al., 2003). Exposure at lOOng/L EE2 also affected reproductive endpoints;
the number of eggs spawned per female over a one-week period was significantly lower,
with large spawning events (100 or more eggs per tank) occurring half as frequently at
this treatment as compared to Ong/L and fertilization success also decreased at this
concentration. The consistency of endocrine effects at lOOng/L and the strong
depressions in spawning and fertilization at the same concentration show that there may
be a link between short-term physiological impairment and long-term reproductive
problems.
In the early life stage bioassay, eggs that had been spawned and fertilized by adult
fish during the fourth week of adult EE2 exposure were exposed at the same
concentrations for 63 weeks (fertilization to 61 weeks post-hatch; Chapter 3). As with
the parental generation, minimal impact of exposure was seen at environmentally-
relevant EE2 exposure levels (<1 Ong/L). Exposure to the highest concentration of EE2
tested (lOOng/L) produced embryos that were larger at hatch and which hatched almost a
day earlier. It is possible that smaller embryos and those that would have hatched later
did not survive exposure to higher concentrations of EE2, resulting in a shift in these
endpoints. As the juvenile fish matured there were no differences in growth or
abnormalities, although survival was higher at lOOng/L EE2. Higher mortality rates at
EE2 concentrations <1 Ong/L compared to juvenile mortality at lOOng/L may be
attributable to increased embryo mortality in the lOOng/L treatment prior to hatch.
Weaker offspring in the lOOng/L treatment may have died prior to hatch, while weaker
offspring at lower concentrations survived hatching and died later.
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When mummichog were exposed to 100ng/L EE2 during early development,
fewer eggs hatched, VTG was induced in females, GSI was reduced in males and the sex
ratio was skewed to >80% female (gonad assessment) and >90% female (secondary sex
characteristics). As fewer individuals hatch and most of them develop into phenotypic
females, these effects have the potential to impact the population. The threshold level for
these types of effects in mummichog appear to lie between 10 and 100ng/L EE2.
Freshwater fish such as the fathead minnow (Ankley et al., 2001; Länge et al.,
2001), Japanese medaka (Seki et al., 2003; Balch et al., 2004), and zebrafish (Olsson et
al., 2003) have been used for development of partial and full lifecycle bioassays for use
in endocrine disruption studies. In comparison, responses of saltwater species to EDS
exposure have been neglected (Leung et al., 2001; MacLatchy et al., 2003). Partial
lifecycle tests have been developed for the estuarine sheepshead minnow (Folmar et al.,
2000; Zillioux et al., 2001; Karels et al., 2003) and mummichog (MacLatchy et al., 2003;
Boudreau et al., 2004; Sharpe et al., 2004). This study begins to address the gap in the
literature regarding the impact of EDS exposure during the life cycle of saltwater or
estuarine fish by linking effects seen in short-term bioassays to long-term impacts.
Short-term exposures with mummichog to EDSs have demonstrated that this
species is sensitive to (anti-)estrogenic and (anti-)androgenic compounds as adults
(MacLatchy et al., 2003; Sharpe et al., 2004) and that early-life development does not
appear to be sensitive at environmentally-relevant concentrations (Boudreau et al., 2004;
Boudreau et al., in press). Three generations of mummichog chronically exposed to
PCBs or mercury via contaminated food showed that effects on sex ratios, the ability to
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reproduce, and male survival were greater in later generations (Matta et al., 2001). This
indicates that impacts may compound with time.
The only consistent negative effects of estrogen exposure in either part of the
study occurred at lOOng/L EE2. VTG induction in male and/or juvenile fish has been
used to indicate estrogen exposure (Denslow et al., 1999; Folmar et al., 2000; MacLatchy
et al., 2003). In this study VTG was present in the plasma of adult male fish and elevated
in the livers of juvenile female fish exposed to lOOng/L. Liver VTG in juvenile males
was not significantly higher, though the level was twice that of male fish not exposed to
EE2 . It is possible that a low sample size (n=5) is driving this (lack of) result and that
VTG was induced in juvenile male fish. Unlike previous short-term EE2 exposures
(MacLatchy et al., 2003), there was no apparent pattern to plasma VTG levels in adult
female fish. This result, along with alterations of spawning output, changes in GSI
during spawning, and modifications of steroid production, suggest that exposure to EE2
may induce gonadal development and/or affect natural reproductive cycling.
During the breeding season (April to September), mummichog GSI naturally
peaks prior to each spawning event. These events occur during the full and new moons
each month and the peak GSI decreases with each successive spawn (Hsiao and Meier,
1989). GSI was found to decrease between weeks 52 and 61 in male and female fish
exposed to 0, 0.1, and lng/L EE2 and in males exposed to lOng/L. This suggests that
these juveniles, though not reproductively active, were following the annual reproductive
cycle, as week 61 coincided with the full moon in August and week 52 fell directly
between the new and full moons of June/July, and one would expect higher GSI near the
full moon had the fish been spawning. Furthermore, females exposed to 10 and lOOng/L
Page 140
did not have a significant decrease in GSI, indicating again that exposure to EE2 can
potentially cause mummichog to deviate from the natural lunar or annual spawning cycle.
Whether prolonged exposure to EE2 can alter female or juvenile reproductive cycling
warrants further investigation. Daily sampling of female mummichog exposed to EE2 for
VTG, steroid hormones and/or GSI for an entire lunar cycle would determine the effects
of EE2 on reproductive cycling.
Changes in circulating steroid hormone levels are a whole organism response to
EDS exposure. In this study, circulating hormone responses were similar to those in
previous 7- and 15- day bioassays (MacLatchy et al., 2003), demonstrating consistency
between studies. However, circulating hormones were not in parallel with in vitro
hormone production or VTG induction in females in this study. Further investigation is
required to determine whether these relationships exist or are inconsistent in
mummichog, and as well the relative sensitivity of these endpoints to EDS exposure
during a lunar reproductive cycle. Also, differences in steroid levels seen between fish
sampled on Day 21 and Day 28 may be attributable to fish being held in single-sex tanks
on Day 21 and mixed sexes on Day 28; this requires further investigation.
Exposure to lOOng/L EE2 affected the GSI of both adult and juvenile male
mummichog. Adult males showed increased GSI when exposed to EE2 and juveniles
showed decreased GSI. Structural changes in normally recrudescing adult males after a
short exposure are unlikely, however, alterations of sperm maturation and/or retention of
viable sperm could produce this result (Gill et al., 2002; Schultz et al., 2003). Juvenile
males may be showing a reduced GSI because of a predominance of immature sperm
and/or testes tissue (Van den Belt et al., 2002). Without gonad histology on both adult
Page 141
and juvenile males, it is impossible to determine the composition of the testes. Gonads
from the juvenile males were saved for future histological analysis.
The total impact on the population of chronic EE2 exposure can be approximated
by a combined assessment of spawning, fertilization, hatching, survival and sex ratios. In
this study, spawning was reduced at lOOng/L EE2, as was subsequent fertilization. Of the
fertilized eggs, fewer hatched when exposed to lOOng/L EE2 . Based upon the spawning,
fertilization, and hatching rates found in this study at Ong/L and lOOng/L, the total
number of larvae available to mature when exposed to lOOng/L EE2 is only about 40% of
the larvae resulting from no EE2 exposure. Once hatched, larvae exposed at lOOng/L had
better survival than all other treatments. It is likely that lower density in these tanks
during the initial 10 weeks contributed to this difference (Boudreau et al., 2004), and this
is not purely a treatment effect. However, the reason for the low survival rates in this
study (14-26% survival from yolk sac absorption) requires investigation, as other studies
with mummichog hatched in the laboratory have seen survival rates as high as 100% at
33 weeks post-hatch (Gutjahr-Gobell, 1998; Boudreau et al., 2004). Rearing conditions,
in particular food quality should be assessed. Breeding trials crossing unexposed females
with exposed males and vice versa will determine whether fertilization difficulties were
attributable to one or both sexes.
High rates of feminization were seen in fish exposed to lOOng/L EE2 . Below this
concentration, sex ratios fluctuated around 50% female, and on each of the three days that
sex ratios were assessed, the sex ratio in fish exposed to lOOng/L EE2 was skewed to
>90% female based upon secondary characteristics, and >80% female based upon gonad
assessment at 4X magnification. Furthermore, based on this study, between 6 and 15%
Page 142
of male fish display female secondary sex characteristics at this level of exposure.
Histological assessment of gonad tissue would clarify this result, and indicate the
proportion of fish with intersex (if any) that was not obvious at 4X magnification. A
breeding trial of mummichog exposed to EE2 from hatch would provide the best
indication of impact on future reproductive potential and population-level impacts.
This life-cycle bioassay investigated whether effects caused by EDS exposure
seen in short-term bioassays may be linked to reproductive potential and progeny impacts
which may translate to population effects in the future. Physiological impairment in
mummichog has been demonstrated at lOOng/L in short-term exposures (this study;
MacLatchy et al., 2003; Sharpe et al., 2004). Reproductive difficulties and subsequent
effects on number of offspring surviving to reproductive maturity and sex ratios at
maturity may be linked to these short-term impairments as they occur at the same
concentration. The possible influence of salinity on fish response to waterborne EE2 also
requires further study. Further investigations as described above, will strengthen this
relationship as similar physiological, biochemical, reproductive and developmental
endpoints are assessed in each generation, and mechanistic explanations are determined.
Exposure to EE2 at concentrations within the range found in the Canadian
environment did not affect mummichog in either generation. This research did identify
endpoints in the life cycle of mummichog that are sensitive to estrogen exposure. Effects
on these endpoints, including circulating levels and production of reproductive steroids,
fecundity, fertility, hatching success, and sex ratios, have the potential to impact long
term population sustainability. Furthermore, this work furthers our ability to extrapolate
between short-term endocrine response and population-level responses in EDS studies.
Page 143
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estrogenic and antiestrogenic chemicals on sheepshead minnows (Cyprinodon
variegatus). Environ. Toxicol. Chem. 22, 855-865.
Länge, R., Hutchinson, T.H., Croudace, C.P., Siegmund, F., Schweinfurth, H., Hampe,
P., Panter, G.H., Sumpter, J.P., 2001. Effects of the synthetic estrogen 17a-
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ethinylestradiol on the life-cycle of the fathead minnow (Pimephales promelas).
Environ. Toxicol. Chem. 20,1216-1227.
Leung, K.M.Y., Morritt, D., Wheeler, J.R., Whitehorse, P., Sorokin, N., Toy, R., Holt,
M., Crane, M. 2001. Can saltwater toxicity be predicted from freshwater data? Mar
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end points in the estuarine mummichog (Fundulus heteroclitus). Environ. Toxicol.
Chem. 22, 996-1008.
Matta, M.B., Linse, J., Caimcross, C., Francendese, L., Kocan, R.M., 2001.
Reproductive and transgenerational effects of methylmercury or Aroclor 1268 on
Fundulus heteroclitus. Environ. Toxicol. Chem. 20, 327-335.
Olsson, P.E., Westerlund, L., Teh, S.J., Billsson, K., Berg A.H., Tysklind, M., Nilsson, J.,
Eriksson, L.-O., Hinton D.E., 1999. Effects of maternal exposure to estrogen and PCB on
different life stages of zebrafish, Danio rerio. Ambio. 28,100-106.
Seki, M., Yokota, H., Matsubara, H., Maeda, M., Tadokoro, H., Kobayashi, K., 2003.
Fish full life-cycle testing for the weak estrogen 4-tert-pentylphenol on medaka (Oryzias
latipes). Environ. Toxicol. Chem. 22,1487-1496.
Page 146
Sharpe R.L., MacLatchy D.L., Courtenay S.C., Van Der Kraak, G.J., 2004. Effects of a
model androgen (methyl testosterone) and a model anti-androgen (cyproterone acetate)
on reproductive endocrine endpoints in a short-term adult mummichog (Fundulus
heteroclitus) bioassay. Aquat. Toxicol. 67, 203-215.
Schultz, I.R., Skillman, A., Nicolas, J.-M., Cyr, D.G., Nagler, J.J., 2003. Short-term
exposure to 17a-ethynylestradiol decreases the fertility of sexually maturing male
rainbow trout (Oncorhynchus mykiss). Environ. Toxicol. Chem. 22, 1272-1280.
Van den Belt, K., Wester, P.W., Van der Ven, L.T.M., Vergeyen, R., Witters, H., 2002.
Effects of ethynylestradiol on the reproductive physiology in zebrafish (Danio rerio):
Time dependency and reversibility. Environ. Toxicol. Chem. 21, 767-775.
Zillioux, E.J., Johnson, I.C., Kiparissis, Y., Metcalfe, C.D., Wheat, J.V., Ward, S.G., Liu,
H., 2001. The sheepshead minnow as an in vivo model for endocrine disruption in
marine teleosts: a partial life-cycle test with 17a-ethynylestradiol. Environ. Toxicol.
Chem. 20, 1968-1978.
Page 147
Candidate’s full name: Rebecca Emily McLeod Peters
Universities attended: Queen’s University (1995-2000), B.ScH, B.Com
Articles in Press:
MacLatchy, D.L., Gormley, K.L., Ibey, R.E.M., Sharpe, R.L., Shaughnessy, K.S., Courtenay S.C., Dube, M.G., Van der Kraak, G.J., 2004. A short-term mummichog (Fundulus heteroclitus) bioassay to assess endocrine response to hormone-active compounds and mixtures. In: Ostrander, G.K. (Ed.) Techniques in Aquatic Toxicology, Vol 2. CRC Press, New York, in press.
Articles Submitted to Refereed Journals:
Peters, R.E.M., Courtenay, S.C., MacLatchy, D.L., 2004. Effects on reproductivepotential and endocrine status in the mummichog (Fundulus heteroclitus) after exposure to 17a-ethynylestradiol in a short-term reproductive bioassay. Aquatic Toxicology, submitted August 2004, 44pp.
Peters, R.E.M., Courtenay, S.C., MacLatchy, D.L., 2004. Effects of 17a-ethynylestradiol on early-life development, sex differentiation and vitellogenin induction in mummichog (Fundulus heteroclitus). Ecotoxicology and Environmental Safety, submitted December 2004, 44pp.
Conference Presentations:
Ibey, R.E.M., Ings, J.S., Courtenay, S.C., MacLatchy, D.L., 2004. Comparison of shortterm and multi-generational bioassays to assess effects of hormonally-active contaminants in mummichog (Fundulus heteroclitus). Oral presentation at Canadian Society of Zoologists (CSZ) 44th Annual Meeting, WolfVille, NS,May 2004.
Ibey, R.E.M., Ings, J.S., Courtenay, S.C., MacLatchy, D.L., 2004. Development of embryonic and juvenile life stages of mummichog (Fundulus heteroclitus) exposed to ethynylestradiol. Poster presentation at Canadian Society of Zoologists (CSZ) 44th Annual Meeting, WolfVille, NS, May 2004.
Page 148
Ibey, R.E.M., Ings, J.S., Courtenay, S.C., MacLatchy, D.L., 2004. Development of lifecycle bioassays with mummichogs to assess environmental effects. Oral presentation at Environmental Effects Monitoring Symposium, Fredericton, NB, February 2004.
MacLatchy, D.L., Ibey, R.E.M., Rickwood, C.J., Dubé, M.G., Hewitt, M.L., 2004.Validation and standardization of freshwater and estuarine fish bioassays to assess reproductive effects of pulp mill effluents. Oral presentation at Environmental Effects Monitoring Symposium, Fredericton, NB, February 2004.
MacLatchy, D.L., Ibey, R.E.M., Ings, J.S., Courtenay, S.C., Van Der Kraak, G.J., 2003. Development of partial and full lifecycle fish bioassays to test for hormonally- active contaminants. Poster presentation at SETAC 24th Annual Meeting, Texas, USA, November 2003.
Ibey, R.E.M., Ings, J.S., Courtenay, S.C., MacLatchy, D.L., 2003. Development of a full life-cycle bioassay for the mummichog (Fundulus heteroclitus). Oral presentation at Canadian Society of Zoologists (CSZ) 43rd Annual Meeting, Waterloo, ON, May 2003.
Publications and presentations prior to October 2004 were under pre-married name,
Rebecca E.M. Ibey.
