Document 6426066

Environmental Toxicology and Chemistry, Vol. 20, No. 10, pp. 2370–2376, 2001
q 2001 SETAC
Printed in the USA
0730-7268/01 $9.00 1 .00
ALTERED DISTRIBUTION OF LIPID-SOLUBLE ANTIOXIDANT VITAMINS IN
JUVENILE STURGEON EXPOSED TO WATERBORNE ETHYNYLESTRADIOL
VINCE P. PALACE,*†‡ ROBERT E. EVANS,† KERRY WAUTIER,§ CHRISTOPHER L. BARON,† JULIETA WERNER,‡
JACK F. KLAVERKAMP,†‡ KAREN A. KIDD,†‡ and TERRY A. DICK‡
†Department of Fisheries and Oceans, Central and Arctic Region, Winnipeg R3T 2N6, Canada
‡University of Manitoba, Department of Zoology, Winnipeg R3T 2N2, Canada
§Biotech West Consulting, Winnipeg R3V 1M4, Canada
( Received 20 October 2000; Accepted 14 March 2001)
Abstract—Studies in mammals have shown that exposure to estrogenic compounds can affect lipid metabolism and plasma
concentrations of lipid-soluble vitamins. However, the potential for estrogenic contaminants to induce these effects in fish has not
yet been examined. The ability of the estrogen analog ethynylestradiol (EE2) to alter concentrations of the lipid-soluble vitamins
A and E in plasma, liver, and kidney was investigated in juvenile lake sturgeon (430 6 20 g). The EE2 was delivered to the
sturgeon in the tank water at nominal concentrations of 0 (control), 15, 60, or 125 ng EE2/L for a period of 25 d. Concentrations
of the egg yolk precursor, vitellogenin, increased dose dependently in plasma. Plasma vitamin E (tocopherol), A1 (retinol), and A2
(dehydroretinol) were elevated by the two highest EE2 treatments compared with the controls. Esterified storage forms of vitamin
A were marginally lower in the livers of fish from the highest EE2 dose group, but vitamin E levels in the liver were not significantly
affected. Concentrations of vitamin E and vitamin A were significantly depleted in the kidney of fish from the two highest EE2
dose groups. Total lipid content was elevated in the gonad of fish treated with the highest dose of EE2 compared with the controls.
Altered lipid and vitamin distribution may be induced by estrogen to facilitate gonadal maturation in sturgeon. Results from these
studies indicate that an examination of the implications for vitamin depletion by estrogenic contaminants in juvenile fish is warranted.
Keywords—Ethynylestradiol
Vitamins
Sturgeon
Lipids
Vitellogenin
penser fulvescens) were exposed to waterborne ethynylestradiol (EE2) at measured concentrations of 0 (control), 15, 60,
or 125 ng/L. Concentrations of the yolk precursor, vitellogenin,
were measured in plasma, and vitamins A and E were evaluated
in plasma, liver, and kidney after 25 d of exposure to EE2.
Evaluation of total lipid in gonad and histological examination
of liver and gonad for evidence of cell damage were also
included in the investigation. Results from this experiment
indicate that exposure to estrogenic compounds can alter lipid
and vitamin distributions in fish.
INTRODUCTION
Concern regarding contaminants that have the potential to
disrupt endocrine functions in aquatic organisms has increased
markedly over the last two decades [1]. While several classes
of contaminants have been identified, those that are capable
of mimicking the effects of the female reproductive hormone,
estrogen, are by far the most widely studied [1]. Estrogenic
contaminants have been identified in a wide array of aquatic
environments [2–6]. Their abilities to disrupt endocrine function and the reproductive physiology of fish have been well
documented and an abundance of biochemical, behavioral, and
physiological surrogate measures of exposure are currently
available to toxicologists. For example, induction of the egg
yolk protein precursor vitellogenin [7–9] and related markers
[10], alterations to gonadal somatic indices [11] and gonad
structure [12,13], reduced steroid hormone production [14,15],
production of tumors [16], effects to sperm motility and egg
production [11,17], and behavioral alterations [18,19] have all
been described in fish exposed to environmental estrogens.
In mammals, estrogenic compounds are known to alter lipid
metabolism and whole-body distribution of the fat-soluble vitamins A and E [20,21]. Examinations of these effects have
not yet been made in fish. However, since maternally derived
vitamins are essential to the development of embryonic fish
[22], an examination of the possibility for altered vitamin metabolism in fish exposed to estrogenic contaminants is warranted.
To examine the effects of exposure to estrogenic compounds on lipid-soluble vitamins, juvenile lake sturgeon (Aci-
MATERIALS AND METHODS
Fish
Juvenile lake sturgeon (one year of age, mean body weight
6 standard error of means [SEM] 5 430 6 20 g) were cultured
at the University of Manitoba, Canada Department of Zoology
laboratories as previously described [23]. Fish were randomly
distributed, seven per tank, into 200-L fiberglass tanks and
acclimated for 10 d. Each tank received 1 L/min of aerated
and dechlorinated Winnipeg, Canada, city tap water. Temperatures were maintained between 11.5 and 13.18C, and dissolved oxygen was at least 90% saturation at all times. During
the acclimation and dosing periods of the experiment, fish were
fed a diet of commercial dry pellet feed at a ration of 1% body
weight/d.
Ethynylestradiol dosing and analysis
The EE2 was purchased from Sigma Chemical Company
(St. Louis, MO, USA) and used directly to make stock solutions for dosing. Four separate stock solutions were prepared
by dissolving an appropriate amount of EE2 in 2 L of absolute
* To whom correspondence may be addressed
(palacev@dfo-mpo.gc.ca).
2370
Lipid-soluble vitamins in ethynylestradiol-exposed sturgeon
ethanol in glass containers. After thorough mixing of the ethanol suspension, 18 L of distilled deionized water was added
for a final stock solution volume of 20 L. A solvent control
solution was prepared consisting of 2 L of ethanol and 18 L
of distilled deionized water. All solutions were maintained at
room temperature and protected from light for the duration of
the experiment.
Following the acclimation period, 200 ml of EE2 stock
solution was introduced into each tank to immediately bring
it to the appropriate concentration of EE2 in ethanol or ethanol
alone, in the case of the control group. Precalibrated peristaltic
pumps were simultaneously started to deliver 1 ml/min of stock
solution into each tank to maintain appropriate EE2-ethanol
target concentrations (0, 20, 100, and 200 ng/L) continuously
for 25 d. Water samples were analyzed weekly to determine
the actual concentration of EE2 in each tank using standard
radioimmunoassay techniques [24] developed by J. Parrott and
S. Brown (National Water Research Institute, Environment
Canada, Burlington, ON). Briefly, a known quantity of an internal standard, 17b-estradiol, and 2% high-performance liquid
chromotography (HPLC) grade methanol were added to a sample of tank water that had been filtered through a 0.45-mm
glass microfiber (GFC) filter. The internal standard and 2%
methanol were used to determine extraction efficiency and
improve estrogen extraction, respectively. The sample was
then drawn by vacuum through Supelclean LC18 filters (Supelco, Bellefonte, PA, USA) that had been conditioned with
15 ml of methanol followed by 10 ml of distilled deionized
water. Columns were stored at 2908C until analysis. After
thawing, the columns were washed with 15 ml of 35% methanol followed by elution of EE2 and the internal standard with
15 ml of 100% methanol. The eluent was dried and resuspended in 1.5 ml of radioimmunoassay (RIA) buffer for use
in the RIA. To determine EE2 in plasma, samples were spiked
with internal standard and extracted with 3:2 ethyl acetate:
hexane (v/v). The ethyl acetate:hexane layer was dried completely under vacuum and the residue was resuspended in RIA
buffer to determine 17b-estradiol and EE2 using the RIA. In
each instance, recovery of the internal standard was used to
calculate an extraction efficiency to correct the EE2 concentrations recovered in each sample. Recoveries of spiked MilliQ water averaged 85% for 1 to 400 ng of EE2.
Tissue analysis
After the exposure period, fish were anesthetized individually by immersion in pH buffered tricaine methanesulfonate
(MS222) 0.8 g/L, pH 7.0. When all fin movement had ceased
(,3 min), fish were removed from the anesthetic, blotted dry,
weighed, and measured. Blood was obtained from the caudal
vein with a preheparinized syringe (50,000 U/ml) and centrifuged at 3,000 g to obtain plasma, which was frozen at 2908C
and protected from light until analysis. Liver and gonad were
dissected and weighed to obtain liver and gonadal somatic
indexes. A subsample of these tissues, as well as posterior
kidney, were fixed in Bouin’s solution for later histological
analysis. The rest of the tissues were frozen in sterile plastic
bags and stored at 2908C for later analysis.
After 24 h in Bouin’s solution, the liver, posterior kidney,
and gonad were washed in several changes of 70% ethanol for
3 d and then embedded in paraffin. Liver sections were cut at
6 mm and kidney and gonad at 8 mm. Tissue sections were
affixed to glass slides and stained with Harris’ hematoxylin
and eosin [25]. To measure the size of hepatocyte nuclei, mi-
Environ. Toxicol. Chem. 20, 2001
2371
croscopic images of the tissues were projected onto a
SummaSketch III digitizer and measured using SigmaScan/
Image, version 1.2. The mean diameter of hepatocyte nuclei
were determined for each fish by measuring the diameter of
40 spherical nuclei, 10 at each of four randomly selected sites.
The mean area of the nuclei (HNA) was calculated using the
formula A 5 pr2, and the hepatocyte volume index was determined by counting the number of nuclei in four randomly
selected tissue areas of 9,000 mm2.
Concentrations of vitellogenin (VTG) were determined in
plasma using a competitive enzyme-linked immunosorbent assay (ELISA). Briefly, 96-well microplates were precoated with
reagent 3.5 mg/ml VTG in 50 mM carbonate buffer, pH 5 9.6,
followed by blocking of the unbound sites in each plate with
5% normal goat serum in the same buffer. Sturgeon plasma
samples that had been appropriately diluted, as well as standards, were incubated for 1 h at 378C with a primary monoclonal antibody against gulf sturgeon VTG (Acipenser oxyrhynchus destoi) at a concentration of 1:6,000 in the incubation
solution of phosphate ELISA buffer, pH 5 7.3. Samples and
standards were then pipetted into the appropriate precoated
wells and incubated for 1 h at 378C. Following this initial
incubation, the secondary antibody was added at a concentration of 1:10,000 and another 1-h incubation at 378C occurred.
The 3,39,5,59-tetramethylbenzidine substrate (Kirkegaard and
Perry Laboratories, Gaithersburg, MA, USA) was then added
to each well to react with the peroxidase enzyme conjugated
to the secondary antibody. After 5 min, the reaction was
stopped by the addition of 1 M phosphoric acid and the yellow
color was read at 450 nm using a Cambridge Model 750 microplate reader (Cambride Technology, Watertown, MA,
USA). Nonspecific binding of the antibodies was quantified
and corrected for in each plate. Reagent VTG used for the
assay was purified from the plasma of vitellogenic lake sturgeon by the Molecular Biomarkers Core Facility, University
of Florida, as described by Denslow et al. [26]. The same
facility also supplied the monoclonal antibody for gulf sturgeon and verified its cross-reactivity with lake sturgeon VTG
[26].
Vitamins E and A were determined in tissues and plasma
using the reverse phase HPLC method of Palace and Brown
[27] with 1% propionic acid added to the mobile phase of 70:
20:10 acetonitrile:dichlormethane:methanol (v/v/v) to improve
separation of closely eluting retinoids. Authentic vitamins for
constructing standard curves were purchased from Sigma with
the exception of vitamin A esters, which were synthesized as
previously described [28]. All solvents were HPLC grade and
the mobile phase was continuously degassed with helium at
30 ml/min in a 4-L vessel. Mobile phase was delivered at 1
ml/min through the Adsorbosphere HS C18 column (Alltech,
Deerfield, IL, USA [250 3 4.6 mm, 5-mm pore size]). Detection was accomplished with a Waters Model 996 (Waters, Milford, MA, USA) diode array detector monitoring absorbance
at 2-nm intervals between 190 and 600 nm. Extraction of the
recordings from the appropriate wavelengths (292 for tocopherol, 325 for retinoids, and 470 for carotenoids), integration,
and data analysis were performed using Millenium32 Software
version 3.05 (Waters Corporations, Milford, MA, USA).
Total lipids in liver and gonad were estimated gravimetrically by freeze drying and pulverizing a sample of tissue. The
fine powder was extracted with 3:2 chloroform:methanol (v/
v) and centrifuged at 3,000 g. An aliquot of chloroform:methanol containing the extracted lipids was evaporated to complete
2372
Environ. Toxicol. Chem. 20, 2001
V.P. Palace et al.
Table 1. Biochemical and physiological responses in juvenile sturgeon exposed to water-borne EE2a
Group
Control
Low dose
Medium dose
High dose
Actual EE2
exposure
(ng/L)
Plasma EE2
(ng/ml)
Plasma VTG
(mg/ml)
0
14
60
123
,0.11
,0.11
1.10 6 0.13b
2.26 6 0.20c
0.4 6 0.4
3.6 6 1.0b
2,902 6 460c
8,071 6 1,272c
LSI
2.33
2.54
2.92
2.89
6
6
6
6
GSI
0.12
0.08
0.10b
0.08b
1.06
1.21
1.12
1.30
6
6
6
6
0.08
0.07
0.10
0.12b
Condition
factor
0.75
0.70
0.75
0.76
6
6
6
6
0.01
0.02
0.02
0.01
HVI
28.8
23.8
19.9
19.9
6
6
6
6
1.6
1.0b
1.2b
0.8c
Liver glycogen
(mg/g)
54.7
49.7
46.0
35.7
6
6
6
6
4.7
7.1
3.1
5.9b
EE2 5 ethynylestradiol; liver somatic index (LSI) 5 liver weight/(body weight 2 liver weight) 3 100; gonadal somatic index (GSI) 5 gonad
weight/(body weight 2 gonad weight) 3 100; condition factor 5 body weight (g)/length (cm)3 3 100; hepatocyte volume index (HVI) 5
number of nuclei per 9,000-mm2 cell area.
b Significantly different from the control group based on ANOVA and Bonferroni’s multiple comparison test (p , 0.05).
c Significantly different from the controls and other dose groups.
a
dryness at 508C in a preweighed aluminum dish. Lipids were
estimated by subtracting the weight of the dish from the total
weight after drying.
Data analysis
All data are presented as mean 6 SEM. Group means were
evaluated using one-way analysis of variance (ANOVA) followed by Bonferroni’s multiple comparison test. Statistical
significance was accepted at the p , 0.05 level. Condition
factor was calculated by the formula CF 5 (wt in g)/(length
in cm)3 3 100 while liver somatic index and gonadal somatic
indexes were derived by gonadal or liver somatic index 5
((organ weight)/(body weight 2 organ weight)) 3 100. For
purposes of calculating gonadal somatic indexes, the tube of
adipose tissue surrounding the gonad proper was included in
the total gonad weight. Since EE2 treatment tended to increase
the percentage of lipids in liver tissue (data not shown), vitamin
concentrations in liver have been expressed relative to the total
amount of lipids. The same effect was not evident in kidney
tissue, and vitamins in this tissue are expressed relative to the
wet weight of tissue.
RESULTS AND DISCUSSION
EE2 exposure
Target water concentrations for the exposures in these experiments were intended to be 20, 100, and 200 ng EE2/L, but
a consistent 30 to 40% loss of EE2 occurred in each exposure
system such that the actual mean exposures were 14, 60, and
123 ng/L (Table 1). The EE2 losses are likely the results of
photodegradation (unpublished observation), microbial degradation [29], and/or partitioning to organic particulate [30],
including mucus that is increasingly sloughed from fish exposed to EE2/ethanol solutions (unpublished observation).
These concentrations are environmentally relevant based on
the sum of EE2 quantities and other estrogenic compounds
recently detected in the direct effluent of Canadian sewage
treatment discharges [6] and in similar international surface
waters [5,7].
Concentrations of EE2 were not detectable in the plasma
of fish from the control or lowest dose groups, but both higher
dose groups had elevated EE2 in plasma (Table 1). While there
is considerable data to relate the exposure levels of EE2 required to induce estrogenic effects, a paucity of information
exists to relate exposure doses to actual levels of EE2 accumulated by fish. Larsson et al. [7] caged juvenile lake trout
downstream from a sewage treatment plant and found EE2
concentrations in the bile of the fish that were 80,000 to
250,000 times greater than those measured in water after two-
or four-week exposures, respectively. In this experiment, concentrations of EE2 were 18-fold greater in the plasma of sturgeon than in the water in both the medium- and high-dose
groups (Table 1). Even this modest level of bioconcentration
is significant considering the potency of EE2 as an estrogenic
compound [2], the low concentrations of hormones required
to elicit physiological responses [1], and the prevalence of
EE2 in surface waters that has already been noted.
Fish condition and histology
Liver somatic index was higher in the medium- and highdose groups compared with the controls and low-dose group
(Table 1). This increase entirely reflects an increase in the
weight of the liver rather than compromised somatic growth
since somatic growth was not different over the course of the
experiment between any of the treatments (data not shown).
Condition factor was monitored as a general indicator of energy reserve for this short duration experiment but was not
significantly different between any of the dose groups (Table
1). Histological evaluation of the liver tissue revealed enlargement of hepatocytes but not nuclei in fish from the two highest
dose groups, as indicated by a decrease in the hepatocyte volume index (Table 1). Hepatocyte volume index is an expression
of the number of nuclei per unit of tissue area [31]. Liver
enlargement and an increase in liver cell size following exposure to EE2 is consistent with previous results in other fish
species [4,7] and likely results from induced synthesis of the
primary egg yolk protein precursor vitellogenin within liver
cells of the exposed sturgeon [32]. The increase of cytoplasmic
basophilia and in cell area occupied by vacuoles in liver tissue
from the exposed sturgeon further supports this conclusion
(Fig. 1) [33].
The VTG increased in the plasma of sturgeon from the low,
medium, and high doses by approximately 7-, 6,000-, and
17,000-fold, respectively, when compared with the control
group (Table 1). Although induction of the VTG response in
fish exposed to waterborne estrogenic contaminants is a function of both exposure time and dose, results from this study
are similar to reported values for salmonid fish species. Rainbow trout (Oncorhynchus mykiss) exposed to 1 ng EE2/L had
up to 1.15 mg VTG/ml in plasma after 10 d of exposure [2].
Mean concentrations of 1.5 mg VTG/ml were also found in
plasma of rainbow trout (Oncorhynchus mykiss) caged for two
weeks downstream from sewage effluent containing several
estrogenic contaminants, including 4.5 ng EE2/L [7].
The gonads of sturgeon differ structurally from the typical
gonads of teleost fish species in that the actual gonad tissue
is embedded in a thick tube of adipose tissue (Fig. 2). Prior
Lipid-soluble vitamins in ethynylestradiol-exposed sturgeon
Fig. 1. Histological sections of liver from a control (A) sturgeon and
a sturgeon exposed to 123 ng/L of waterborne ethynylestradiol for 25
d (B). Magnification, 3460.
to gonad maturation, which in female sturgeon does not occur
until 14 to 23 years of age, there is an increase in the amount
of adipose tissue surrounding the gonad (T.A. Dick, unpublished observation). In fact, during vitellogenesis, an increase
in the plasma concentrations of other lipoproteins [34] may
facilitate this transfer of lipids to the gonads of female sturgeon. Using gravimetric extraction, we found that total lipids
increased in the gonads and surrounding adipose tissue, but
only in sturgeon exposed to the highest dose of EE2 were these
increases statistically significant (Fig. 3). It should be noted
that the early developmental stage of the gonads from these
juvenile sturgeon (;one year) made determining sex-specific
effects of EE2 on the gonads impossible as evidenced by the
undifferentiated primordial germ cells of the gonads (Fig. 2,
panel B).
The formation of VTG may also divert energy from vital
proteins and lipids in order to support gonad growth [4]. Livers
from the high-dose group appeared to contain less glycogen
than the control group, based on the loss of typical moth-eaten
cellular architecture that is normally attributed to glycogen
(Fig. 1) [35]. Biochemical analysis of the glycogen content in
livers of fish from this experiment confirmed a significant depletion of liver glycogen at the highest EE2 exposure (Table
1).
Lipid-soluble vitamins
Vitamin depletion has been shown to occur in response to
a variety of organic contaminants present in aquatic environ-
Environ. Toxicol. Chem. 20, 2001
2373
Fig. 2. Histological sections of the gonad of a control group sturgeon
showing the gonad embedded in a tube of adipose tissue (A) (magnification, 332) and a primordial germ cell at higher magnification
(3525) (B).
ments. For example, vitamins A and E are depleted in fish
exposed to planar organochlorines [27,36,37]. With respect to
sturgeon, those captured from systems heavily contaminated
with organochlorines had lower tissue concentrations of vi-
Fig. 3. Total lipid in the adipose tube surrounding the gonads of control
and ethynylestradiol exposed sturgeon. Data are presented as mean
6 standard error of means (SEM), n 5 10. *, significantly different
from the control group based on analysis of variance (ANOVA) followed by Bonferroni’s multiple comparison test (p , 0.05).
2374
Environ. Toxicol. Chem. 20, 2001
Fig. 4. Vitamins E, A1, and A2 in plasma of control and ethynylestradiol-exposed sturgeon. Data are presented as mean 6 standard error
of means (SEM), n 5 10. *, significantly different from the control
group based on analysis of variance (ANOVA) followed by Bonferroni’s multiple comparison test (p , 0.05); **, significantly different
from other ethynylestradiol dose groups.
tamin A [38,39]. However, we previously reported that vitamin
A was not depleted in sturgeon orally exposed to 2,3,7,8tetrachlorodibenzofuran and that vitamin E was only depleted
in the liver, kidney, and plasma of these fish at the highest
dose (1.6 mg/g) and exposure duration (27 d) [23]. Depletion
of vitamins in organisms exposed to planar organochlorines
has been postulated to result from induced metabolic enzyme
V.P. Palace et al.
activity [37,40], altered transport capacity, or inhibited feeding
[41].
Information linking estrogenic contaminant exposure to lipid-soluble vitamin depletion in fish is lacking. However, some
studies in mammals have shown that vitamin homeostasis can
be altered by exposure to synthetic estrogens. Plasma concentrations of vitamin A are typically elevated in women prescribed oral contraceptives that contain ethynylestradiol [21].
Vitamin E concentrations can also be altered in plasma and
liver tissue following EE2 exposure, likely as a result of increased production of oxygen radicals inducing lipid peroxidation and the subsequent consumption of vitamin E [42,43].
In fish, stored vitamins are mobilized to the gonads for incorporation into oocytes during vitellogenesis [38], an event
that is normally mediated by estrogen but that can also be
induced by estrogenic contaminants [1].
Like many other fish, sturgeon utilize two forms of vitamin
A, i.e., vitamin A1 (retinol) and vitamin A2 (dehydroretinol).
The A2 isoform has lower biological activity based on results
obtained in mammals but usually predominates in sturgeon in
both plasma and as stored fatty acid esters in tissues [38,39].
Plasma concentrations of both forms of vitamin A were elevated in sturgeon from the two highest dose groups in this
experiment (Fig. 4). Similarly, concentrations of vitamin E
were significantly higher in the plasma of sturgeon from the
two highest dose groups compared with the control fish (Fig.
4). Since plasma concentrations of vitamins are derived largely
from visceral stores [44], it is likely that long-term increases
in circulating levels would be at the expense of vitamins in
these organs. The short duration of these experiments, however, resulted in only marginally lower concentrations (p 5
0.08) of stored vitamin A esters in the livers of EE2-exposed
fish (Fig. 5). The free alcohol forms of vitamin A1 and A2
and vitamin E in liver were not significantly affected by EE2
exposure over the course of this experiment (data not shown).
In contrast with the results from the liver, vitamin concentrations were significantly affected in the kidneys of sturgeon
exposed to EE2. Vitamin E and A1 were significantly lower
Fig. 5. Vitamin E and total vitamin A ester concentrations in liver of control and ethynylestradiol-exposed sturgeon. Data are presented as mean
6 standard error of means (SEM), n 5 10. ;, marginally different from the control group based on analysis of variance (ANOVA) followed
by Bonferroni’s multiple comparison test (p 5 0.08).
Lipid-soluble vitamins in ethynylestradiol-exposed sturgeon
Fig. 6. Vitamins E, A1, and A2 in kidney of control and ethynylestradiol-exposed sturgeon. Data are presented as mean 6 standard error
of means (SEM), n 5 10. *, significantly different from the control
group based on analysis of variance (ANOVA) followed by Bonferroni’s multiple comparison test (p , 0.05); **, significantly different
from other ethynylestradiol dose groups.
in the two highest dose groups and vitamin A2 was depleted
only at the highest EE2 exposure (Fig. 6). Whereas fatty acid
esters of vitamin A1 and A2 are routinely detected in kidneys
of many fishes [27], these compounds were not detected in
the kidneys of sturgeon from any of the dose groups in this
experiment.
Several mechanisms might contribute to the altered vitamin
concentrations observed in EE2-exposed sturgeon from these
studies. First, oxidative stress induced by exposure to EE2
could increase vitamin metabolism and contribute to visceral
organ depletion of these compounds. It has been shown that
EE2 promotes oxidative stress through its conversion to catechol estrogen metabolites that undergo redox cycling to liberate free radicals [45,46]. Vitamins E and A are both physiologically important antioxidants and can be depleted from
visceral organs by oxidative stress events [47]. Administration
of either of the vitamins has been shown to reduce the oxidative
damage induced in livers of rats exposed to EE2 [43].
On first consideration, it appears that oxidative stress arising from EE2 exposure cannot explain elevated concentrations
of vitamins in the plasma of EE2-exposed fish that occur concurrent with declining storage organ levels. However, oxidative
stress events have also been shown to increase the activity of
retinol ester hydrolase enzymes that are responsible for hydrolyzing stored vitamin A esters in the liver of mammals so
that the vitamin is mobilized from the liver and into the plasma
[47]. Whether this mechanism operates after EE2 exposure in
fish remains to be investigated.
Other mechanisms described in mammals may be responsible for the altered vitamin and lipid distribution following
EE2 exposure. For example, studies have shown that EE2 induces cholestasis and therefore reduced clearance rates of compounds normally excreted in bile [48]. Biliary excretion is an
important aspect of the normal turnover of vitamins [44]. It
has also been shown that EE2 can also alter the activities of
lipase enzymes in the liver and adipose tissue, leading to dis-
Environ. Toxicol. Chem. 20, 2001
2375
rupted lipid profiles in the plasma [20]. Different lipid concentrations can, in turn, influence the plasma complement of
vitamins, which are derived from the liver and other major
storage sites.
This study has shown for the first time that ethynylestradiol,
an estrogenic contaminant of concern in Canadian freshwaters,
can alter the distribution of the lipid-soluble vitamins A and
E in exposed fish. These short-term experiments raise the question of whether prolonged exposure to EE2 can result in more
significantly depleted visceral stores of lipid-soluble vitamins.
In previous experiments, fish exposed to vitamin-depleting
contaminants did not exhibit depleted visceral storage organ
concentrations until several weeks after the initial exposures
[36]. Considering the importance of vitamins A and E for
supporting normal embryonic development in fish, the potential for estrogenic contaminants to alter vitamin metabolism
and embryonic development following long-term exposure
warrants further investigation. Studies designed to examine
the mechanisms for altered vitamin distribution in fish exposed
to estrogenic contaminants are currently ongoing in our laboratory.
Acknowledgement—These studies were supported by grants from
Health Canada’s Toxic Substances Research Initiative, the Department
of Fisheries and Oceans’ Environmental Science Strategic Research
Fund, and the Chemical Manufacturers Association. T.A. Dick was
supported by research grants from National Science and Engineering
Research Council and Manitoba Hydro. The authors acknowledge the
technical assistance of Mingchau Lu as well as J. Parrott and S. Brown
of the National Water Research Institute in Burlington Ontario for
their work in developing the RIA assay for measuring EE2 in water
and plasma samples.
REFERENCES
1. Kime DE. 1998. Endocrine Disruption in Fish. Kluwer, Norwell,
MA, USA.
2. Purdom CE, Hardiman PA, Bye VJ, Eno NC, Tyler CR, Sumpter
JP. 1994. Estrogenic effects of effluents from sewage treatment
works. Chem Ecol 8:275–285.
3. Routledge EJ, Sheahan D, Desbrow C, Brighty GC, Waldock M,
Sumpter JP. 1998. Identification of estrogenic chemicals in STW
effluent 2. In vivo responses in trout and roach. Environ Sci
Technol 32:1559–1565.
4. Harries JE, Sheahan DA, Jobling S, Matthiessen P, Neall P, Sumpter JP, Tylor T, Zaman N. 1997. Estrogenic activity in five United
Kingdom rivers detected by measurement of vitellogenesis in
caged male trout. Environ Toxicol Chem 16:534–542.
5. Belfroid AC, Van der Horst A, Vethaak AD, Schafer AJ, Rijs
GBJ, Wegener J, Cofino WP. 1999. Analysis and occurrence of
estrogenic hormones and their glucuronides in surface water and
waste water in the Netherlands. Sci Total Environ 225:101–108.
6. Ternes TA, Stumpf M, Mueller J, Haberer K, Wilken RD, Servos
M. 1999. Behaviour and occurrence of estrogens in municipal
sewage treatment plants. I. Investigations in Germany, Canada
and Brazil. Sci Total Environ 225:81–90.
7. Larsson DGJ, Adelofsson-Erici M, Parkkonen J, Petterson M,
Berg AH, Olsson PE, Forlin L. 1999. Ethynylestradiol—An undesired fish contraceptive? Aquat Toxicol 45:91–97.
8. Carlson DB, Williams DE. 1999. Sex-specific vitellogenin production in immature rainbow trout. Environ Toxicol Chem 18:
2361–2363.
9. Sumpter JP, Jobling S. 1995. Vitellogenesis as a biomarker for
estrogenic contamination of the aquatic environment. Environ
Health Perspect 103(Suppl. 7):173–178.
10. Kramer VJ, Miles-Richardson S, Pierens SL, Giesy JP. 1998.
Reproductive impairment and induction of alkaline-labile phosphate, a biomarker of estrogen exposure, in fathead minnows
(Pimepheles promelas) exposed to waterborne 17b-estradiol.
Aquat Toxicol 40:335–360.
11. Jobling S, Sheahan D, Osborne JA, Matthiessen P, Sumpter JP.
1996. Inhibition of testicular growth in rainbow trout (Oncor-
2376
12.
13.
14.
15.
16.
17.
18.
19.
20.
21.
22.
23.
24.
25.
26.
27.
28.
29.
Environ. Toxicol. Chem. 20, 2001
hynchus mykiss) exposed to estrogenic alkylphenolic chemicals.
Environ Toxicol Chem 15:194–202.
Gimeno S, Komen H, Vanderbosch P, Bowmer T. 1997. Disruption
of sexual differentiation in genetic male common carp (Cyprinus
carpio) exposed to alkylphenol during different life stages. Environ Sci Technol 31:2884–2890.
Gray MA, Metcalf CD. 1997. Induction of testis–ova in Japanese
medaka (Oryzias latipes) exposed to p-nonlyphenol. Environ
Toxicol Chem 16:1082–1086.
MacLatchy DL, Van der Kraak G. 1995. The phytoestrogen bsitosterol alters the reproductive endocrine status of goldfish. Toxicol Appl Pharmacol 134:305–312.
Alfonso LOB, Campbell PM, Iwama GK, Devlin RH, Donaldson
EM. 1997. The effect of the aromatase inhibitor fadrazole and
two polynuclear aromatic hydrocarbons on sex steroid secretion
by ovarian follicles of coho salmon. Gen Comp Endocrinol 106:
169–174.
Cooke JB, Hinton DE. 1999. Promotion by 17b-estradiol and bhexachlorocyclohexane of hepatocellular tumours in medaka,
Oryzias latipes. Aquat Toxicol 45:127–145.
Goodman LR, Hansen DJ, Manning CS, Faas LF. 1982. Effects
of kepone on the sheepshead minnow in an entire life-cycle toxicity test. Arch Environ Contam Toxicol 11:335–342.
Matthiessen P, Logan JWM. 1984. Low concentration effects of
endosulfan insecticide on reproductive behaviour in the tropical
cyclid fish Sarotherodon mossambicus. Bull Environ Contam
Toxicol 33:575–583.
Holdway DA, Dixon DG. 1986. Impact of pulse exposure to
methoxychlor exposure on flagfish (Jordanella floridae) over one
reproductive cycle. Can J Fish Aquat Sci 43:1410–1415.
Demacker PN, Staels B, Stalenhoef AF, Auwerx J. 1991. Increased removal of beta-very low density lipoproteins after ethynyl estradiol is associated with increased mRNA levels for hepatic
lipase, lipoprotein lipase, and the low density lipoprotein receptor
in Watanabe heritable hyperlipidemic rabbits. Arterioscler
Thromb 11:1652–2659.
Mooij PN, Thomas CM, Doesburg WH, Eskes TK. 1991. Multivitamin supplementation in oral contraceptive users. Contraception 44:277–288.
Cahu CL, Cuzon G, Quazuguel P. 1995. Effect of highly saturated
fatty acids, alpha-tocopherol and ascorbic acid in broodstock diet
on egg composition and development of Penaeus indicus. Comp
Biochem Physiol 112A:417–424.
Palace VP, Dick TA, Brown SB, Baron CL, Klaverkamp JF. 1996.
Oxidative stress in lake sturgeon (Acipenser fulvescens) orally
exposed to 2,3,7,8-tetrachlordibenzofuran. Aquat Toxicol 35:79–
92.
McMaster ME, Munkittrick KR, Van Der Kraak GJ. 1992. Protocol for measuring circulating levels of gonadal steroids in fish.
Can Tech Rep Fisheries Aquat Sci 1836.
Edwards JE. 1967. Methods for the demonstration of intercellular
substances of the connective tissues. In McClung Jones R, ed,
McClung’s Handbook of Microscopical Technique. Hafner, New
York, NY, USA.
Denslow ND, Chow M, Chow MM, Bonomelli S, Folmar LC,
Heppell SA, Sullivan CV. 1997. Development of biomarkers for
environmental contaminants affecting fish. In Rolland RM, Gilbertson M, Peterson RE, eds, Chemically Induced Alterations in
Functional Development and Reproduction of Fishes. SETAC,
Pensacola, FL, USA, pp 73–86.
Palace VP, Brown SB. 1994. HPLC determination of tocopherol,
retinol, dehydroretinol and retinyl palmitate in tissues of lake char
(Salvelinus namaycush) exposed to coplanar 3,39,4,49,5-pentachlorobiphenyl. Environ Toxicol Chem 13:473–476.
Palace VP, Hill MF, Farahmand F, Singal PK. 1999. Mobilization
of antioxidant vitamin pools and hemodynamic function after
myocardial infarction. Circulation 99:121–126.
Layton AC, Gregory BW, Seward JR, Schultz TW, Sayler GS.
2000. Mineralization of steroidal hormones by biosolids in waste-
V.P. Palace et al.
30.
31.
32.
33.
34.
35.
36.
37.
38.
39.
40.
41.
42.
43.
44.
45.
46.
47.
48.
water treatment systems in Tennessee USA. Environ Sci Technol
34:3925–3931.
Lai KM, Johnson KL, Scrimshaw MD, Lester JN. 2000. Binding
of waterborne steroid estrogens to solid phases in river and estuarine systems. Environ Sci Technol 34:3890–3894.
Leatherland JF, Sonstegard RA. 1984. Pathobiological responses
of feral teleosts to environmental stressors: Interlake studies of
the physiology of Great Lakes salmon. In Cairns VM, Hodson
PV, Nriagu JO, eds, Contaminants Effects on Fisheries. John
Wiley, New York, NY, U.S.A. pp 115–150.
Takashima F, Hibiya T. 1995. An Atlas of Fish Histology: Normal
and Pathological Features, 2nd ed. Kodansha, Tokyo, Japan.
Gimeno S, Komen H, Jobling S, Sumpter J, Bowmer T. 1998.
Demasculinisation of sexually mature male common carp, Cyprinus carpio, exposed to 4-tert-pentylphenol during spermatogenesis. Aquat Toxicol 43:93–109.
Tyler CR, Sumpter JP. 1996. Oocyte growth and development in
teleosts. Rev Fish Biol Fisheries 6:287–318.
Bucher F, Hofer R, Salvenmoser W. 1992. Effects of treated papermill effluents on hepatic morphology in male bullhead (Cottus
gobio L.). Arch Environ Contam Toxicol 23:410–419.
Palace VP, Brown SB, Metner DA, Lockhart WL, Muir DCG,
Klaverkamp JF. 1996. Mixed function oxidase enzyme activity
and oxidative stress in lake trout (Salvelinus namaycush) exposed
to 3,394,49,5-pentachlorobiphenyl. Environ Toxicol Chem 15:
955–960.
Branchaud A, Gendron A, Fortin R, Anderson PD, Spear PA.
1995. Vitamin A stores, teratogenesis, and EROD activity in white
sucker, Catostomus commersoni, from Riviere des Prairies near
Montreal and a reference site. Can J Fish Aqua Sci 52:1703–
1713.
Doyon C, Boileau S, Fortin R, Spear PA. 1998. Rapid HPLC
analysis of retinoids and dehydroretinoids stored in fish liver:
Comparison of two lake sturgeon populations. J Fish Biol 53:
973–986.
Ndayibagira A, Cloutier MJ, Anderson PD, Spear PA. 1995. Effects of 3,39,4,49-tetrachlorobiphenyl on the dynamics of vitamin
A in brook trout (Salvelinus fontinalis) and intestinal retinoid
concentrations in lake sturgeon (Acipenser fulvescens). Can J
Fish Aquat Sci 52:512–520.
Palace VP, Klaverkamp JF, Baron CL, Brown SB. 1997. Metabolism of 3H-retinol by lake trout (Salvelinus namaycush). Aquat
Toxicol 39:321–332.
Spear PA, Higueret P, Garcin H. 1994. Effects of fasting and
3,39,4,49,5,59-hexabromobiphenyl on plasma transport of thyroxine and retinol. J Toxicol Environ Health 42:173–183.
Ciavatti M, Blache D, Renaud S. 1989. Hormonal contraceptive
increases plasma lipid peroxides in female rats. Relationship to
platelet aggregation and lipid biosynthesis. Arteriosclerosis 9:84–
89.
Ogawa T, Higashi S, Kawarada Y, Mizumoto R. 1995. Role of
reactive oxygen in synthetic estrogen induction of hepatocellular
carcinomas in rats and preventative effect of vitamins. Carcinogenesis 16:831–836.
Blumhoff R, Green MH, Berg T, Norum KR. 1990. Transport and
storage of vitamin A. Science 250:399–404.
Chen J, Gokhale M, Li Y, Trush MA, Yager JD. 1998. Enhanced
levels of several mitochondrial mRNA transcripts and mitochondrial superoxide production during ethynyl estradiol-induced hepatocarcinogenesis and after estrogen treatment of HepG2 cells.
Carcinogenesis 19:2187–2193.
Barth A, Landmann G, Liepold K, Zapf H, Muller D, Karge E,
Klinger W. 1999. Influence of oestrogens on formation of reactive
oxygen species in liver microsomes of differently aged male Wistar rats. Exp Toxicol Pathol 51:282–288.
Palace VP, Hill MF, Khaper N, Singal PK. 1999. Metabolism of
vitamin A in the heart increases after a myocardial infarction.
Free Rad Biol Med 26:1501–1507.
Bouchard G, Yousef IM, Tuchweber B. 1994. Decreased biliary
glutathione content is responsible for the decline in bile saltindependent flow induced by ethynyl estradiol in rats. Toxicol
Lett 74:221–233.