PNAS
Collapse
of
a
fish
population
after
exposure
to
a
synthetic
estrogen
Karen
A.
Kidd*,, Paul
J. Blanchfield*,
Kenneth H. Mills*, Vince P. Palace*,
Robert E. Evans*,
James
M. Lazorchak,
and
Robert W. Flick
http://www.pnas.org/cgi/content/full/104/21/8897
*Fisheries
and
Oceans
Canada,
Freshwater
Institute, 501 University Crescent, Winnipeg, Manitoba,
Canada
R3T 2N6; and
Molecular
Indicators
Research
Branch,
United States
Environmental
Protection Agency,
26 West Martin
Luther King Drive, Cincinnati,
OH 45268
Edited by Deborah
Swackhamer,
University of
Minnesota,
Minneapolis,
MN, and
accepted
by the Editorial
Board
March
29, 2007 (received for review October
27, 2006)
Abstract
Municipal
wastewaters
are
a
complex mixture containing
estrogens
and
estrogen
mimics that
are
known to
affect
the reproductive health
of
wild fishes.
Male
fishes
downstream
of
some wastewater
outfalls
produce vitellogenin (VTG) (a
protein normally
synthesized by females
during oocyte maturation)
and
early-stage
eggs in their testes, and
this feminization
has
been attributed
to
the presence of
estrogenic
substances
such as
natural
estrogens
[estrone or 17
-estradiol
(E2)], the synthetic
estrogen
used in birth-control pills [17
-ethynylestradiol
(EE2)], or weaker
estrogen
mimics such as
nonylphenol in the water.
Despite widespread
evidence that
male
fishes
are
being feminized, it is not known whether these low-level, chronic exposures
adversely
impact
the sustainability
of
wild populations.
We conducted a
7-year,
whole-lake
experiment at
the Experimental
Lakes
Area
(ELA)
in northwestern Ontario,
Canada,
and
showed that
chronic exposure
of
fathead
minnow (Pimephales
promelas)
to
low concentrations
(56 ng·L1) of
the potent 17-ethynylestradiol
led to
feminization
of
males
through the production of
vitellogenin mRNA
and
protein, impacts
on gonadal
development as
evidenced by intersex in males
and
altered
oogenesis in females,
and,
ultimately,
a
near
extinction of
this species from the lake.
Our observations
demonstrate
that
the concentrations
of
estrogens
and
their mimics observed in freshwaters
can
impact
the sustainability
of
wild fish
populations.
endocrine disrupters | fathead
minnow | municipal
wastewaters
| population
| level effects | whole-lake
experiment
There is considerable
evidence that
fishes
inhabiting
waters
that
receive untreated
municipal
wastewaters
or effluents from municipal
wastewater
treatment
plants
(MWTPs) are
exposed to
chemicals
that
affect
reproductive endocrine function. Male
fish
downstream
of
some wastewater
outfalls
produce vitellogenin (VTG) mRNA
and
protein, associated
with oocyte maturation
in females,
and
early-stage
eggs in their testes (13).
This feminization
has
been linked to
the presence of
estrogenic
substances
such as
the natural
estrogen
17-estradiol
(E2) and
the synthetic
estrogen
17-ethynylestradiol
(EE2) (4). Natural
and
synthetic
estrogens
are
not completely broken down in current MWTP processes (5),
and,
as
a
result, are
discharged
into
receiving waters
in both treated
and
untreated
wastewaters
and
found in the aquatic
environment at
low parts
per trillion concentrations
(typically
<5 ng·L1; refs. 68).
These effluents contain
mixtures of
individual
estrogens
and
their mimics that
differ in their ability
to
elicit estrogenic
responses (e.g., ref. 9). For this reason,
total
estrogenicity
(expressed as
E2 equivalents)
of
an
effluent or water
sample
is determined either by summing concentrations
of
individual
compounds after
adjusting
these concentrations
by the compound's estrogenic
potency (relative
to
E2; ref. 10) or with bioassays
(e.g., refs. 1012).
Using these approaches,
total
E2 equivalents
of
up to
147 and
17 ng·L1 have
been measured
in final
effluents and
surface
waters,
respectively (1012).
Within this group of
substances,
the estrogen
used in birth-control pills, EE2, is one of
the more potent estrogens
present (9) and
has
been linked to
the feminization
of
male
fishes
in rivers receiving municipal
wastewater
(4, 6).
Despite growing documentation
on the feminization
of
male
fishes
in waterways
receiving municipal
effluents, a
critical
question in the field of
endocrine disruption research
is whether these low-level, chronic exposures
adversely
impact
wild populations
(13). Although
laboratory
studies have
shown decreased
reproductive success of
fish
exposed to
<15 ng·L1 of
EE2 (14, 15), it is unknown whether this
response would be observed in wild populations
and
whether it would result in a
subsequent decline in abundances.
To
assess
the ecological
risk posed by this class
of
compounds, we must understand
population-level
effects of
estrogens
and
their mimics on aquatic
organisms.
The fathead
minnow (Pimephales
promelas)
is a
common species in North America,
and
its range
extends from the southern United States
to
northern Canada
(16). It is an
important
food source for numerous game
fish
species, such as
lake
trout (Salvelinus
namaycush),
walleye
(Sander
vitreus), and
northern pike (Esox lucius). In the lakes
used for this study, fathead
minnow have
a
lifespan
of
4
years,
but few individuals
live past
2 years
of
age
(17). Asynchronous
spawning
starts
in early
summer and
extends for a
period of
2
months; multiple females
will typically
spawn
in the nest of
a
single male,
who will care
for developing eggs until hatching.
Sexual
maturity
occurs during the second year
of
life, so most fish
spawn
for only one season.
Therefore, 2 consecutive years
of
reproductive failure
cause
catastrophic
declines in abundance
(18, 19). Fathead
minnow have
been widely adopted
by the scientific community for toxicity
testing because
of
their widespread
distribution, ease
of
laboratory
culture, well characterized
biology and
life history,
and
the large
family
(Cyprinidae)
of
fishes
that
they represent (2022).
It is a
freshwater
equivalent
of
the "miner's canary."
We describe the results of
a
7-year,
whole-lake
study at
the Experimental
Lakes
Area
(ELA)
in northwestern Ontario,
Canada,
to
assess
the subcellular-level
through population-level
effects of
the potent synthetic
estrogen
EE2 on fathead
minnow. The concentrations
of
EE2 achieved
in the experimental
lake,
Lake
260, during the 3 years
of
additions
were within the range
of
those observed in untreated
and
treated
municipal
wastewaters
(5, 6) and
below the total
E2 equivalents
(with EE2 having
an
E2 equivalent
of
0.191.9; ref. 23) measured
in effluents and
receiving waters
(1012).
Results and
Discussion
Measures
of
VTG mRNA
and
protein are
used to
assess
exposure
of
fishes
to
estrogens
and
estrogen
mimics, and
elevated
concentrations
in males
are
indicative
of
the presence of
these substances
in effluents or surface
waters
(1, 3). We used these endpoints to
assess
the subcellular
responses of
fathead
minnow from Lake
260 to
EE2 additions.
Seven weeks after
the first estrogen
additions
to
Lake
260 began
in 2001, VTG concentrations
in male
and
female
fathead
minnow were elevated
when compared
with preaddition
or reference lake
data
(Fig. 1). Whole-body homogenates
of
males
from Lake
260 had
concentrations
of
VTG that
were three orders of
magnitude
greater
than
reference samples,
and
this response was
sustained
in each
of
the 3 years
of
EE2 additions.
Female
fathead
minnow also
produced more VTG after
exposure
to
EE2 when compared
with reference samples
(reference lake
fish
had
2.5% of
the VTG concentrations
observed in treated
fish
from Lake
260; Fig. 1), although
their response was
less dramatic
than
that
observed for male
fish.
Liver VTG mRNA
was
measured
in samples
collected from the EE2-amended
and
reference lakes
concurrently with those used for VTG protein analyses.
Results from these mRNA
analyses
showed patterns
similar
to
the VTG protein expression. Over the 3 years
of
dosing, male
fathead
minnow from Lake
260 had
mean
± SD normalized
liver VTG mRNA
values
ranging
from 0.422 ± 0.685 to
1.22 ± 0.181. Liver VTG mRNA
values
for males
from the two reference lakes
were on average
<0.11.6% of
those for the EE2-exposed males.
In addition,
males
from Lake
260 had
VTG mRNA
values
(0.904 ± 0.437) that
were more than
an
order of
magnitude
higher than
female
fish
(0.045 ± 0.048) collected from the reference lakes
during the same
time periods. Females
collected from Lake
260 in the spring and
mid-summer of
each
year
of
EE2 additions
had
gene expression that
ranged
from 0.834 ± 0.879 to
0.378 ± 0.053, and
this response was
82152% of
the values
in females
from the reference lakes.
These among-lake
differences were highest in the fall;
for example,
females
collected in 2002 from Lake
260 exhibited levels of
VTG mRNA
expression (0.849 ± 0.793) that
were more than
an
order of
magnitude
higher than
those of
reference females
(0.045 ± 0.048). This trend was
also
observed in the accompanying
protein data,
demonstrating
that
EE2 amendments
prolonged the expression of
VTG mRNA
and
protein in males
and
females
well past
the breeding season.
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Fig.
1. Mean
± SE (n = 47) VTG concentrations
in whole-body homogenates
of
male
(Lower) and
female
(Upper) fathead
minnow captured
in 19992003 from reference Lakes
114 and
442 and
from Lake
260 before and
during additions
of
56 ng·L1 of
EE2 (low catches
of
fish
in Lake
260 in 2004 and
2005 did not allow
for these analyses
in the latter
2 years
of
the study). |
Wild fishes
exposed to
municipal
wastewater
effluents and
fishes
exposed to
estrogens
in laboratory
studies have
shown disruption in gonadal
development as
evidenced by the presence of
both male
and
female
gonadal
tissues (i.e., intersex) and
decreased
gonadosomatic
indices (GSIs) (2, 14, 15).
In this study, testicular
tissues of
all
of
the male
fathead
minnow collected during the first spring after
EE2 additions
displayed
delayed
spermatogenesis,
widespread
fibrosis, and
malformations
of
the tubules (see ref. 24). Testicular
germinal
tissue from all
EE2-exposed males
consisted primarily
of
spermatogonia
instead
of
the spermatocytes
that
were normally
observed at
this time of
year
in the reference lakes
and
in Lake
260 before EE2 additions.
In the spring of
2002, the GSI for males
from Lake
260 averaged
0.40 ± 0.21% (n = 10). This mean
was
well below the GSI values
of
1.39 ± 0.38% (n = 15) and
2.27 ± 0.41% (n = 10) recorded for males
from reference Lakes
114 and
442, respectively. This arrested
testicular
development continued in 2003 and
2004, and
four of
nine males
captured
in the spring of
2003 had
ova-testes
with the presence of
primary-stage
oocytes (Fig. 2). It has
been shown previously that
male
fish
with intersex have
reduced fertilization
success because
of
decreased
sperm production and
mobility (25).
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Fig.
2. Gonadal
tissue sections from fathead
minnow sampled
in early
May
2003. (A)
Ovary
from reference Lake
442 with small,
dark,
primary-stage
oocytes situated
between cortical
alveolar-stage
oocytes and
large
vitellogenic oocytes. (B) Ovary
from EE2-amended
Lake
260 demonstrating
an
atretic
follicle (arrow).
(C) Testis from reference Lake
442. (D) Testis from EE2-amended
Lake
260 demonstrating
intersex; arrows
indicate
primary-stage
oocytes among
the remnants
of
testicular
tissue. (Scale
bars:
A,
B, 300 ΅m; C, D, 80 ΅m.) |
High VTG production in female
fathead
minnow from Lake
260 was
followed by delayed
ovarian
development in individuals
collected the next spring (Fig. 2) (ref. 24),
which is consistent with observations
for the other dominant
minnow species in this lake,
the pearl
dace
(Margariscus
margarita)
(26). Oocytes in female
fathead
minnow collected from Lake
260 the first spring after
EE2 amendments
began
were at
a
much earlier
stage
of
development (cortical
alveolar
versus vitellogenic) than
those from fathead
minnow in nearby
pristine lakes
or in Lake
260 before EE2 additions.
However, delayed
oocyte maturation
was
not evident in ensuing spring samples
nor were there any
consistent effects on gonadosomatic
indices during the 3 years
of
EE2 amendments
(data
not shown). Each
year
subsequent to
2002 there was
an
increasing
number of
fish
ovaries
that
contained
a
few atretic
follicles, which were rarely
observed in reference and
preaddition
fish
(Fig. 2).
Index trap
netting [adults
and
young-of-the-year
(YOY)] and
minnow traps
(adults)
were used to
examine
changes
in the fathead
minnow population
size and
structure in the fall
during each
of
the 7 years
in both Lake
260 and
Lake
442 (Fig. 3). Trap-net
catch-per-unit-effort
(CPUE) data
showed that
the mean
abundance
of
this species varied
considerably
(9.6 ± 2.2 to
490 ± 68) in reference Lake
442 from 1999 to
2005 (Fig. 3A).
In contrast,
the CPUE of
fathead
minnow decreased
consistently after
EE2 additions
to
Lake
260 from 180 ± 48 and
50 ± 43 (before additions)
to
0.1 ± 0.01 in the fifth year
after
additions
began
(Fig. 3B). Lake
260 and
Lake
442 fathead
minnow CPUE data
derived from fall
catches
using minnow traps
(data
not shown) showed the same
trends as
the trap-net
data
and,
overall,
the two methods were highly comparable
(Spearman
rank
correlation:
t12 = 6.45, R = 0.88, P < 0.0001). Although
fathead
minnow populations
typically
fluctuate
in size from year
to
year
because
of
variable
recruitment, year
class
failures
are
not observed in unmanipulated
lakes
(17, 27).
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Fig.
3. Length frequency
distributions of
fathead
minnow captured
in trap
nets in reference Lake
442 (A)
and
Lake
260 (B) (amended
with 56 ng·L1 of
EE2 in 20012003) during the fall
of
19992005. Distributions for each
fall
have
been standardized
to
100 trap-net
days.
Mean
± SE daily
trap-net
CPUE data
for adults
and
juveniles for the fall
catches
are
shown in the panels. |
The fathead
minnow population
in Lake
260 collapsed
in the fall
of
2002, after
the second season
of
EE2 additions,
because
of
a
loss of
young-of-the-year
(Fig. 3B). This reproductive failure
was
also
observed in the third season
of
amendments
and
continued for an
additional
2 years
after
the EE2 additions
had
ceased,
although
a
few small
individuals
were caught
each
year,
indicating
some reproduction was
occurring. The loss of
smaller
size classes
of
fathead
minnow was
not observed in reference Lake
442 (Fig. 3A),
a
system with similar
species composition, water
volume, and
trophic status
(24, 28). In addition,
the mean
± SE size of
adult
fish
from Lake
260 increased
from 50 ± 0.28 mm (n = 300) in 2001, just after
the first EE2 additions,
to
62 ± 0.18 mm (n = 263) in the fall
of
2003, because
of
a
shift in the age
structure of
the remaining
population.
Population-level
effects documented in this study mimicked what
was
observed during a
long-term acidification
study at
the Experimental
Lakes
Area
(18). At
lower pH, there was
near
total
extinction of
fathead
minnow and
an
increase
in the median
size of
the remaining
adult
population
(19). The reproductive failure
and
near
extirpation
of
fathead
minnow in Lake
260 has
not yet been observed in the longer-lived pearl
dace
in this system (26), and
this suggests that
life-history
characteristics
such as
lifespan
are
important
determinants
of
a
species' risk from exposure
to
estrogens
and
estrogen
mimics.
The results from this whole-lake
experiment demonstrate
that
continued inputs of
natural
and
synthetic
estrogens
and
estrogen
mimics to
the aquatic
environment in municipal
wastewaters
could decrease
the reproductive success and
sustainability
of
fish
populations.
Chronic exposure
of
fathead
minnow to
a
potent synthetic
estrogen
led to
feminization
of
males
through production of
VTG mRNA
and
protein, continued production of
VTG in females
beyond the normal
breeding season,
impacts
on gonadal
development, as
evidenced by intersex in males
and
altered
oogenesis in females,
and
a
near
extinction of
this species from the lake.
Likely because
of
its short life-cycle, the fathead
minnow was
the first to
show population
collapse
in this experiment. This response implies that
short-lived fish
species may
generally
be at
greatest
risk from exposure
to
estrogens
and
their mimics, but chronic exposure
of
longer-lived species to
these substances
may
result in the loss of
these populations
as
well.
Materials
and
Methods
The experimental
system, Lake
260, is located
in an
undisturbed watershed
at
the Experimental
Lakes
Area,
has
a
surface
area
of
34 ha
and
a
maximum
depth of
14 m, and
contains
naturally
reproducing populations
of
lake
trout, white sucker (Catostomus
commersoni), pearl
dace,
and
fathead
minnow, as
well as
several
other small-bodied
fish
species. We studied Lake
260 and
two nearby
reference systems, Lakes
114 and
442, for 2 years
before EE2 additions,
3 years
during the amendments
to
the experimental
lake,
and
2 years
after
the additions
were discontinued. The physical,
chemical,
and
biological
characteristics
of
Lake
260 and
the reference lakes
are
detailed
in refs. 24 and
28.
We added
EE2 (Schering AG,
Berlin, Germany)
three times weekly to
Lake
260 starting
in May
2001, after
the lake
had
stratified,
and
additions
were continuous during the open-water
seasons
of
20012003 for a
period of
2021 weeks. For each
EE2 addition,
the required mass
[calculated
by using its predetermined half-life
in the water
column (K.A.K.,
unpublished data),
the volume of
the epilimnion, and
the mean
concentration
of
EE2 measured
the previous week] was
dissolved in 30 ml of
DIG-grade
methanol
(Caledon
Laboratories,
Georgetown,
ON, Canada)
and
added
to
the propeller wash
of
the boat
as
it was
driven in transects
across
the lake.
Mean
± SD epilimnetic concentrations
of
EE2 were quantified
weekly by using replicate
samples
collected at
each
of
five sites around
the lake.
EE2 was
extracted
onto
preconditioned C-18 columns (Supelco, Oakville,
ON, Canada),
eluted with 100% DIG methanol,
and
quantified
along
with the internal
standard
testosterone
by using standard
RIA
techniques and
GC-MS confirmation
for 10% of
the samples.
Additional
details
on extraction
methods and
EE2 concentrations
have
been described in refs. 24 and
28. Seasonal
mean
± SD EE2 concentrations
in Lake
260 were 6.1 ± 2.8 (n = 189), 5.0 ± 1.8 (n =
169), and
4.8 ± 1.0 (n = 170) ng·L1 in 20012003,
respectively.
For VTG mRNA
and
protein analyses,
fathead
minnow were collected by using live-minnow traps
in the spring and
fall
of
19992003 and
in the summer of
2000 and
2001 from Lake
260 and
two reference lakes
(n = 410 individuals
per sampling
date
for each
analysis).
Each
fish
was
handled
according
to
Canadian
Animal
Care
protocols
(approved
by the Freshwater
Institute Animal
Care
Committee) and
individually
euthanized
in pH-buffered tricaine
methanesulfonate
(250 mg·L1). For VTG analyses,
whole fish
were placed
in sterile plastic
bags,
immediately
frozen on dry ice, and
kept at
90°C until analyzed.
Whole-body homogenates
were used to
quantify
the VTG by using the indirect competitive ELISA
methods detailed
previously (24). Reagent
VTG from fathead
minnow was
obtained
from the Core Biomarker
Facility
(University of
Florida,
Gainesville,
FL). The primary
antibody
used for the analysis
was
a
mouse anti-carp
VTG monoclonal
antibody
(ND-2D3; Biosense Laboratories,
Bergen, Norway)
raised
against
VTG from common carp
(Cyprinus carpio).
Fish
captured
for VTG mRNA
analysis
were euthanized
and
dissected immediately
after
capture,
and
their livers were placed
in RNAlater
(Ambion,
Austin,
TX) for transport
to
the laboratory.
Liver VTG mRNA
was
measured
by using quantitative
real-time
PCR and
normalized
to
18S ribosomal
gene expression, an
invariant
ribosomal
gene, by using methods detailed
previously (29).
Impacts
on fathead
minnow gonadal
development in Lake
260 were best examined
in the spring of
each
year
when gonads
were mature
and
not affected
by the asynchronous
spawning
season.
Fish
were collected from Lake
260 and
the two reference lakes
each
spring from 1999 to
2003 as
described above;
whole fish
were euthanized,
immersed in Bouin's solution, and
processed by using standard
histological
techniques as
described previously (30).
Index trap
netting (adults
and
young-of-the-year)
and
minnow traps
(adults
only) were used to
examine
changes
in the fathead
minnow population
size and
structure in the fall
during each
of
the 7 years
in both Lake
260 and
reference Lake
442. Each
September, at
least
three trap
nets were set in each
lake
and
fished
from 7 to
33 consecutive days;
nets were emptied every 23 days.
Lake
260 and
Lake
442 CPUE data
(not shown) were also
derived from fall
minnow trapping
(30 traps
set per lake
for 1011 consecutive days
and
emptied daily)
and
compared
with CPUE data
from fall
trap
netting (shown in Fig. 3).
Acknowledgements
We thank
D. W. Schindler, G. T. Ankley,
K. R. Munkittrick, R. Lδnge, and
two anonymous
reviewers for their helpful comments on this manuscript.
Fish
population
sampling
was
conducted by S. M. Chalanchuk
and
L. S. Tate
[Department
of
Fisheries
and
Oceans
(DFO), Winnipeg, MB, Canada].
This research
was
supported by the American
Chemistry Council Long Range
Initiative
Program
(K.A.K.),
the Fisheries
and
Oceans
Canada
(DFO) Environmental
Sciences Strategic
Research
Fund (K.A.K.,
V.P.P., R.E.E., P.J.B., and
K.H.M.), the Canadian
Federal
Toxic
Substances
Research
Initiative
(K.A.K.,
V.P.P., and
R.E.E.), and
the Canada
Research
Chair
program
(K.A.K).
EE2 was
supplied by Schering AG.
VTG analyses
were done by K. Wautier
(DFO, Winnipeg). EE2 analyses
were conducted by L. Vandenbyllaardt
(DFO, Winnipeg) and
K. Londry (University of
Manitoba,
Winnipeg, MB, Canada).
Footnotes
Abbreviations:
CPUE, catch-per-unit-effort;
E2, 17-estradiol;
EE2, 17-ethynylestradiol;
VTG, vitellogenin.
To
whom correspondence should be sent at
the present address:
Canadian
Rivers Institute and
Biology Department,
University of
New Brunswick, Saint
John, New Brunswick, Canada
E2E 4P1. E-mail:
kiddk@unbsj.ca
Freely available
online through the PNAS
open access
option.
Author
contributions: K.A.K.,
P.J.B., K.H.M., and
V.P.P. designed research;
K.A.K.,
P.J.B., K.H.M., V.P.P., R.E.E., and
J.M.L. performed research;
P.J.B., K.H.M., V.P.P., R.E.E., J.M.L., and
R.W.F. analyzed
data;
and
K.A.K.,
P.J.B., K.H.M., V.P.P., R.E.E., J.M.L., and
R.W.F. wrote the paper.
The authors
declare
no conflict of
interest.
This article
is a
PNAS
Direct Submission. D.S. is a
guest editor
invited by the Editorial
Board.
© 2007 by The
National
Academy
of
Sciences of
the USA
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