Uptake, metabolism and excretion of bisphenol A in the rainbow trout (Oncorhynchus mykiss)

Article abstract of DOI:10.1016/S0166-445X(01)00157-6

The uptake, metabolism and excretion of the oestrogenic chemical bisphenol A (BPA) were studied in juvenile rainbow trout (Oncorhynchus mykiss). BPA was detectable in plasma, liver and muscle after 2 h of water exposure at 0.44 μM (100 μg BPA/l), and a st

Full text of DOI:10.1016/S0166-445X(01)00157-6

Aquatic Toxicology 55 (2001) 75–84
www.elsevier.com/locate/aquatox
Uptake, metabolism and excretion of bisphenol A in the
rainbow trout (Oncorhynchus mykiss)
Christian Lindholst, Søren N. Pedersen 1, Poul Bjerregaard *
Institute of Biology, SDU-Odense Uni6ersity, Campus6ej 55, DK-5230 Odense M, Denmark
Received 26 September 2000; received in revised form 22 January 2001; accepted 23 January 2001
Abstract
The uptake, metabolism and excretion of the oestrogenic chemical bisphenol A (BPA) were studied in juvenile
rainbow trout (Oncorhynchus mykiss). BPA was detectable in plasma, liver and muscle after 2 h of water exposure at
0.44 mM (100 mg BPA/l), and a steady state was reached within 12–24 h. The concentration of the glucuronidated
degradation product in the plasma was about twice that of the parent compound. A plasma half life of BPA was
calculated as 3.75 h following injection of the compound. The vitellogenin synthesis was measured in response to the
BPA treatment, and a lag period of 5 and 7 days between injection of the compound and a significant vitellogenin
response was observed for females and males, respectively. At the time of the vitellogenin response no BPA could be
detected in the liver tissue from either male or female fish. These results indicate that fish briefly exposed to elevated
levels of oestrogenic chemicals might develop a response several days later. © 2001 Elsevier Science B.V. All rights
reserved.
Keywords: Toxico-kinetics; Oncorhynchus mykiss; Bisphenol A; Bisphenol A glucuronide; Vitellogenin; Xenoestrogen
and chemical formulations (Liehr et al., 1998).
EDCs reach the aquatic environment mainly
through sewage effluents thereby posing a poten-
tial threat to the organisms living in these ecosys-
tems (Matthiessen and Sumpter, 1998). Field
studies conducted on wild fish from English rivers
show a high incidence of intersex among males
living near sewage effluents (Jobling et al., 1998).
Other experiments show that estrogenic responses
can be detected in fish up to 5 km downstream
from the site of sewage discharge (Harries et al.,
1997). In addition to the estrogenic effects ob-
served in natural aquatic environments, labora-
tory studies have revealed various endocrine
1. Introduction
For the past decades, an increasing effort has
been made to elucidate the endocrine disrupting
properties of natural as well as man-made com-
pounds with hormonal activity (Toppari et al.,
1996; Tyler et al., 1998). The endocrine disrupting
chemicals (EDCs) can be found in a wide variety
of matrices such as foodstuff, packaging materials
* Corresponding author. Tel.: +45-65502456; fax: +45-
65930457.
E-mail address: poul@biology.ou.dk (P. Bjerregaard).
1 Present address: Cheminova A/S, Department of Chemical
Analysis, Postbox 9, DK-7620 Lemvig, Denmark.
0166-445X/01/$ - see front matter © 2001 Elsevier Science B.V. All rights reserved.
PII: S0166-445X(01)00157-6

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C. Lindholst et al. / Aquatic Toxicology 55 (2001) 75–84
The present study examines the toxicokinetic
behaviour of BPA in the rainbow trout
Oncorhynchus mykiss. The uptake, accumulation
and excretion of nonmetabolised BPA was moni-
tored in blood plasma and muscle- and liver
tissue. Furthermore, the metabolism of BPA was
estimated by direct measurements of BPAGA in
the blood plasma. The quantification of degrada-
tion product was carried out by means of an
enzymatically synthesised BPAGA standard. Fi-
nally the yolk protein vitellogenin (VTG) was
measured in the plasma throughout the experi-
mental period, in order to characterise the estro-
genic response associated with BPA exposure.
disrupting properties of the compounds in ques-
tion (Christiansen et al., 1998; Panter et al., 1998).
One of the compounds known to mimic the
actions of natural oestrogen is bisphenol A (BPA)
(Dodds and Lawson, 1936; Benjonathan and
Steinmetz, 1998). In fish, an increased concentra-
tion of the yolk protein vitellogenin (VTG) has
been observed following injection and water expo-
sure of the compound (Christiansen et al., 2000;
Lindholst et al., 2000). Apart from the endocrine
disrupting effects observed for BPA and other
hormonally active compounds, it is necessary to
consider the uptake, metabolism and excretion in
order to fully understand their modes of action. A
rapid excretion or conversion of an oestrogenic
compound into a less potent metabolite, might be
a way in which an organism can reduce its oestro-
genic load. Since EDCs often include phenols and
other aromatic groups (e.g. alkyl phenols, phtha-
lates, parabens and bisphenols) conjugation of
polar moieties by phase II degradation enzymes is
expected. The addition of a glucuronyl or sulfonyl
group to the parent molecule is generally believed
to increase the water solubility thereby facilitating
the subsequent excretion (Giroud et al., 1998).
As regards the disposition of BPA, most of our
present knowledge is based on mammalian stud-
ies. In 1966, Knaak and Sullivan examined the
various routes of excretion of BPA in the rat,
following oral exposure. They concluded that
bisphenol A glucuronic acid (BPAGA) was a
major degradation product subjected to renal ex-
cretion. Within recent years, the specific isoforms
of glucuronosyl- and sulfonyl transferase, respon-
sible for the enzymatic phase II conjugation of
BPA have been characterised in the rat (Yokota
et al., 1999; Suiko et al., 2000). In piscine systems,
knowledge about metabolism and excretion of
BPA, as well as most other EDCs, is scarce.
Larsson et al. (1999) examined the presence of
BPAGA in the bile from caged fish exposed to
sewage effluent. Other studies measured the activ-
ity of phase II degradation enzymes in response to
xenoestrogenic exposure (Andersson et al., 1985;
Arukwe et al., 1997). However, the level of the
BPAGA production as well as the excretion rate
of BPA in fish has until now been unexplored.
2. Materials and methods
2.1. Experimental set-up
Juvenile rainbow trout (O. mykiss: 90–130 g)
were purchased from a local fish farm, and accli-
mated for 10 days in 400-l tanks supplied with
running fresh groundwater at 15°C. Fish were
held in a 12-h light, 12-h dark photoperiod and
were not fed during the experiment.
The experiment was carried out in a flow-
through system consisting of 400-l tanks, fitted
with circulatory pumps (Eheim 1250, Germany).
Administration of water was controlled through
the use of a centrifugal pump (UPS 25-40 Grund-
fos, Denmark) at a flow rate of 960 l/day, whereas
BPA was administered through a peristaltic pump
(Ole Dich Instrumentmakers, Denmark) at a flow
rate of 60 ml/day.
One group containing 54 fish was continuously
exposed to 100 mg BPA/l (Aldrich, Germany)
dissolved in 96% ethanol while a control group,
designated exposure control, containing 36 fish
was administered 96% ethanol. In another group,
54 fish were injected intraperitoneally (i.p.) with a
solution of 48% ethanol containing 35 mg BPA/
ml giving a final tissue concentration of 35 mg
BPA/kg. A group of 36 fish, designated injection
control, was exposed to the vehicle alone.
At 0, 2, 6, 12, 24, 48, 72, 120 and 168 h, six fish
from each of the exposure groups and four from
each control group were anaesthetised in 0.02%

C. Lindholst et al. / Aquatic Toxicology 55 (2001) 75–84
77
phenoxyethanol (Sigma, Germany) and a 1000 ml
blood sample was obtained from the dorsal vein
for vitellogenin, BPA and BPAGA analysis.
Blood was collected into heparinised 1.5 ml Ep-
pendorf tubes and centrifuged for 4 min at
The extracted material from the preceding sam-
ple clean-up procedure, dissolved in 65%
methanol, was diluted into 20% methanol and
made up with 0.2% formic acid before further
purification via HPLC fractionating. Aliquots of 2
ml were applied to a 200×16 mm C-18 column
pre-equilibrated with 0.2% formic acid. BPAGA
or BPAGA-d16 was eluted with a linear gradient
of 0–100% 0.2% formic acid in methanol for 40
min with a flow rate of 3 ml/min. The acquired
HPLC fractions were pooled and evaporated to
dryness by means of a vacuum dryer.
4000×g in
a Dich centrifuge. Plasma was
pipetted into 1 ml Eppendorf tubes, and stored at
−80°C until analysis. Subsequent to the sampling
of blood, the liver was removed and finally a 2 g
muscle sample was obtained from the thoracic
region. All tissue samples were stored at −80°C.
2.2. Preparation of BPAGA and BPAGA-d16
standards
2.4. Characterisation of BPAGA
The purified BPAGA consisted of fine white
needle-shaped crystals with a melting point of
150–154°C. BPAGA was dissolved in deuterated
methanol (CD3OD) (Merck, Germany) and
analysed by NMR spectrometry. The 13C NMR
spectrum (500 MHz) was as follows (numbers of
C-atoms relate to the molecular structure shown
in Fig. 1): 161.1 (C-6), 144.6 (C-7), 144.0 (C-16),
134.8 (C-10), 131.0 (C-13), 116.8 (C-14), 116.6
(C-9), 105.4 (C-15), 103.7 (C-8), 90.6 (C-1), 65.4
(C-5), 64.5 (C-3), 62.6 (C-2), 61.1 (C-4), 30.7
(C-11), 19.5 (C-12). Further structural verification
of the synthesized compound was obtained from a
full scan mass spectrum (m/z=50–450) using the
same instrumental settings as described in Section
2.5, apart from an increased cone voltage of 150
V (Fig. 1).
BPAGA and BPAGA-d16 standards were syn-
thesised using a modified version of Stevenson
and Hubl (1999) based on the enzymatic conjuga-
tion of BPA or BPA-d16 and UDP-b-D-glucuronic
acid (UDPGA) by the liver microsomal enzyme
UDP-glucuronyl transferase (UDPGT). Bovine
liver UDPGT and UDPGA were purchased from
Sigma, Denmark. The incubation buffer consisted
of 50 mM Tris (pH 8.0), 1 mM dithiothreitol, 6
mM CaCl2, 3% (w/v) bovine serum albumin and 8
mM UDPGA. UDPGT was added giving an en-
zyme activity of 0.044 U/ml. The reaction was
started by the addition of BPA or BPA-d16 dis-
solved in the co-solvent acetonitrile giving sub-
strate and co-solvent concentrations of 4 mM and
2% (v/v), respectively. The incubation was carried
out at 35°C for 24 h.
2.5. Chromatographic parameters for BPAGA
and BPAGA-d16
2.3. Purification of BPAGA
Subsequent to incubation the reaction mixture
was centrifuged at 15 000×g for 15 min in a
Sigma 3K30 centrifuge (Struers, Denmark) to re-
move suspended proteins. The supernatant was
diluted with 4 vol. water and pH was adjusted to
2.0 with acetic acid (Merck, Germany). Sample
amounts of 4 ml were applied on a Waters Sep-
Pak C-18 500 mg cartridge previously conditioned
with 5 ml methanol and 5 ml 0.2% formic acid.
Before eluting the sample in 4 ml 65% methanol
the cartridge was washed with 5 ml 0.2% formic
acid and allowed to dry for 3 min.
A Hewlett Packard 1100 LC-MSD equipped
with an electrospray ionisation (ESI) interface was
used for separating, identifying and quantifying
the actual concentrations of BPAGA and
BPAGA-d16 in plasma. Chromatography was per-
formed using a 150×2.1 mm Zorbax Eclipsed
XDB reverse-phase HPLC column (5-mm particle
size) equilibrated with 100% solvent A (0.2% for-
mic acid). BPAGA and BPAGA-d16 were eluted
for 2 min with solvent A followed by a linear
gradient of 0–100% solvent B (100% methanol,
0.2% formic acid) for 20 min. The flow rate was

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C. Lindholst et al. / Aquatic Toxicology 55 (2001) 75–84
0.4 ml/min. The MSD was tuned automatically
using the built-in calibrant delivery system. Se-
lected ion monitoring (SIM) was performed in
negative mode for the ions m/z 403 (BPAGA) and
m/z 417 (BPAGA-d16). The MS conditions were
as follows: drying gas temperature 250°C, drying
gas flow 10 l/min, nebulizing gas pressure 20 psig,
capillary voltage 4000 V, fragmentor voltage 70
V. Peak area ratios of standards and samples with
respect to internal standard were computed using
Hewlett Packard Chemstation software. The con-
centrations of BPAGA in samples were deter-
mined by interpolation from the appropriate
standard curve. The quantification limit for
BPAGA using this method was 50 nM based on a
200-ml plasma sample. All chemicals used in the
extractions and analytical procedures were of ana-
lytical grade.
scribed by Pedersen and Lindholst (1999). The
limit of quantification for water samples was 2.2
nM based on a 20-ml sample whereas in tissue the
limit of quantification was 219 nM (50 ng BPA/g)
based on 1 g of fresh tissue.
2.7. Plasma analysis
Plasma samples were analysed for their content
of BPA and BPAGA. Approximately 200 ml
plasma were dissolved in 4 ml 0.2% formic acid
(pH 2) and internal standards were added. For
the quantification of BPA, 30 ml of a deuterated
(BPA-d16) internal standard was used, whereas 30
ml deuterated BPAGA (BPAGA-d16) was used as
internal standard for BPAGA. The samples were
applied on a Waters Sep-Pak C-18 500 mg car-
tridge conditioned with 5 ml methanol and 5 ml
0.2% formic acid, and desorbed by 4 ml 65%
methanol made up with 0.2% formic acid. Finally,
the samples were evaporated to dryness in a vac-
uum dryer (Maxi dry lyo, Denmark) and redis-
solved in 300 ml 0.2% formic acid in methanol.
Aliquots of 10 ml were used for analysis. The
analysis of BPA was carried out using LC-MS
equipped with an APCI interface as described by
Pedersen and Lindholst (1999).
2.6. Water and tissue analysis
Every second day two 20-ml water samples
were collected from each tank. The water, liver
and muscle samples were analysed for their con-
tent of BPA using microwave assisted solvent
extraction (tissue samples only) and solid phase
extraction followed by LC-MS analysis as de-
Fig. 1. Structure (a) and mass spectrum (b) of bisphenol A glucuronic acid (BPAGA) at a fragmentor voltage of 150 V. Numbers
on the molecular structure of BPAGA refer to the interpretation of the 13C NMR spectrum (see text).

C. Lindholst et al. / Aquatic Toxicology 55 (2001) 75–84
79
Fig. 2. Bisphenol A (BPA) and bisphenol A glucuronic acid (BPAGA) in the blood plasma of rainbow trout (mean9SEM, n=6).
The fish were either water-exposed to 0.44 mM BPA (A) or intraperitoneally injected with 154 mmol BPA/kg (B). The dashed and
solid horizontal lines indicate the limit of quantification for BPA and BPAGA, respectively.
2.8. Vitellogenin analysis
mass spectrum of BPAGA recorded at a fragmen-
tor voltage of 150 V, shows the [M–H]− ion at
m/z=403 and the [M-177]− ion at m/z=227
originating from BPA. The fragments observed at
m/z=175 and 113 probably result from the glu-
curonic acid moiety.
VTG in the plasma samples was measured by a
direct sandwich ELISA based on polyclonal,
affinity purified antibodies, as described in Chris-
tiansen et al. (2000). The present assay had a
quantification limit of 0.04 mg VTG/ml.
3.2. Water, plasma and tissue le6els of BPA and
BPAGA
2.9. Statistical analysis
Results obtained from the ELISA were sub-
jected to one-way analysis of variance (ANOVA)
followed by Tukey’s test for multiple compari-
sons. All statistics were calculated using the soft-
ware program Systat 7.0 with PB0.05 as the level
of significance.
In one of the experimental groups, fish were
exposed to a nominal concentration of 0.44 mM
BPA (100 mg BPA/l). Every second day through-
out the experimental period the actual water con-
tent of BPA was measured showing a mean
concentration of 0.43 mM and a SEM of 0.004.
No BPA could be detected in water from the
injection group or any of the control groups.
Plasma samples, collected at 0, 2, 6, 12, 24, 48,
72, 120 and 168 from each of the two exposure
groups were analysed for their content of BPA
and BPAGA (Fig. 2). In the water-exposed group,
an increase in BPA concentration was observed
within the first 12 h reaching a maximum concen-
tration of 1.5 mM (0.35 mg/ml). The same pattern
3. Results
3.1. Characterisation of BPAGA
Following purification of BPAGA, a characteri-
sation of the generated standard was carried out
by 13C NMR and mass spectrometry (Fig. 1). The

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C. Lindholst et al. / Aquatic Toxicology 55 (2001) 75–84
was seen for the appearance of the degradation
product BPAGA in the plasma; concentrations
reached a steady state in the region of 2.5–3.7
mM, approximately 2–2.5 times higher than the
level of BPA. In the injection group, maximum
concentrations of BPA and BPAGA of 83 and
265 mM were observed after 2 and 6 h, respec-
tively. A decrease in BPA to levels below the limit
of quantification was reached within 72 h. A
first-order exponential decay regression line was
Excretion rates for BPA in plasma, liver and
muscle were calculated from the data obtained in
the injection group (Fig. 4). The BPA levels mea-
sured 2 h after injection were normalised to 100%
giving comparable values on BPA excretion in the
three compartments. The fastest excretion was
observed in the liver with 19.4, 7.7 and 1.5% BPA
remaining after 6, 12 and 24 h, respectively,
whereas 48.9, 14.4 and 2.0% were remaining in the
plasma. The lowest rate of excretion was, how-
ever, found in muscle where 73.8, 26.0 and 1.7%
BPA remained after 6, 12 and 24 h, respectively.
Plasma samples collected from each of the two
control and exposed groups were analysed for
their content of VTG (Fig. 5). The straight lines
shown in Fig. 5A,B represent mean VTG levels of
all control male or female fish sampled through-
out the experiment. The means (9SEM) for
males and females are: 0.132 (90.069) and 0.395
(90.081) mg VTG/ml in the water-exposed group
(Fig. 5A) and 0.108 (90.026) and 0.464 (9
0.136) mg VTG/ml in the injection group (Fig.
5B). A one-way ANOVA followed by Tukey’s test
for multiple comparisons was calculated for male
and female fish from both exposure groups. Sig-
calculated to y=(mean9SEM)e−(mean9SEM)
=
(118.591.7)e−(0.18590.004)t giving a BPA plasma
half life of 3.75 h.
The internal doses of BPA in liver and muscle
were measured throughout the 7 days of experi-
mentation (Fig. 3). The concentration of BPA
increased in liver and muscle to 2.4 and 1.7 mM
after 24 and 48 h, respectively, in water-exposed
fish. A decline at 72 and 120 h was observed in
both liver and muscle to levels between 0.8 and
0.9 mM. In fish injected with BPA, maximum
concentrations of 311 and 83 mM were reached
within 2 h in liver and muscle, respectively. Con-
centrations dropped below the limit of quantifica-
tion in liver after 168 h and in muscle after 120 h.
Fig. 3. Internal concentrations of bisphenol A (BPA) in liver and muscle tissue of rainbow trout (mean9SEM, n=6). The fish were
either water-exposed to 0.44 mM BPA (A) or intraperitoneally injected with 154 mmol BPA/kg (B). The dashed lines indicate the limit
of quantification.

C. Lindholst et al. / Aquatic Toxicology 55 (2001) 75–84
81
When fish were exposed through the water,
absorption of BPA reached a steady state level
within 6–12 h in plasma, liver and muscle (Figs. 2
and 3). This uptake rate is in agreement with
previous results reported for other phenolic com-
pounds, e.g. a steady state was reached in 4 h in
rainbow trout exposed to phenol (McKim et al.,
1999). After reaching a steady state, bioconcentra-
tion factors (BCF) for BPA were calculated to
approximately 3.5–5.5 for plasma, liver and mus-
cle. For plasma, the BCF was calculated as (BPA/
ml plasma)/(BPA/ml water) while BCF for liver
and muscle tissue were calculated as (BPA/g tis-
sue)/(BPA/ml water). We have previously re-
ported BCFs at a BPA concentration of 100 mg/l
to be 10.8 and 2.2 in liver and muscle, respectively
(Lindholst et al., 2000).
The uptake of BPA in the blood plasma of
rainbow trout coincides with the appearance of
the glucuronidated degradation product BPAGA.
Since the BPA molecule contains two phenol moi-
eties, formation of conjugated degradation prod-
ucts such as sulfonated and glucuronidated
metabolites are expected. Apart from the phase II
metabolites generated from BPA, the introduction
of an additional hydroxyl group to the parent
molecule, leading to quinines, is an alternative
degradation pathway observed in mammals
(Atkinson and Roy, 1995). The formation of glu-
curonidated metabolites in response to BPA expo-
sure was first described by Knaak and Sullivan
(1966) in a mammalian system. The results from
the present experiment show that the concentra-
tion of BPAGA increases to levels above BPA
within 2–6 h. This suggests that glucuronidated
products can be used as more sensitive indicators
of xenobiotic exposure than the parent com-
pounds. Previous studies suggest that BPAGA is
accumulated in mg/g quantities in gall fluid from
rainbow trout exposed to BPA (Larsson et al.,
1999). One of the consequences of glucuronida-
tion is a reduction in the partition coefficient
(Kow), meaning that compounds are made more
water soluble and thereby easier to excrete.
Giroud et al. (1998) showed that the introduction
of a glucuronosyl group reduces the Kow by two
orders of magnitude. This would result in a reduc-
tion in the log Kow for BPA from 3.4–3.8 to
Fig. 4. A comparison of the bisphenol A (BPA) excretion from
plasma, muscle- and liver tissue of rainbow trout injected with
154 mmol BPA/kg. The excretion rates are expressed as % BPA
remaining relative to the levels observed at 2 h.
nificant differences in the VTG levels were found
between the control group and female fish sam-
pled at 120 and 168 h together with males sam-
pled at 168 h in the injection study (PB0.05). No
significance was observed between fish from any
of the water-exposed groups.
4. Discussion
The aim of the present study was to examine
the uptake, accumulation, degradation and excre-
tion of a well established xenooestrogen in a
teleostean system. Two separate experiments, one
on the water uptake and another on the excretion
following compound injection were performed us-
ing the rainbow trout (O. mykiss) as a test organ-
ism.
Compound
administration
in
the
water-exposed group was controlled by the use of
a continuous flow-through system, assuring good
agreement between nominal and actual concentra-
tions. In the present study the actual water con-
centration of BPA only deviated by 2.3% from
nominal values over a 7-day period.

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C. Lindholst et al. / Aquatic Toxicology 55 (2001) 75–84
approximately 1.4–1.8 for BPAGA (Staples et al.,
1998). It is known that the activity of degradation
enzymes can vary as a consequence of enzyme
latency (Andersson et al., 1985). The lag period
can either be ascribed to low activity of the exist-
ing enzymes or low levels of available enzyme.
However, the rapid formation of BPAGA in the
water-exposed group suggests that the conjugat-
ing enzyme, UDP glucuronosyl transferase, is al-
ready present in sufficient amounts to convert the
absorbed BPA.
The excretion of BPA from plasma, liver and
muscle was measured following an intraperitoneal
injection of the compound. The fact that maxi-
mum concentrations of BPA are found in the
three compartments 2 h after injection suggests
that the compound is readily absorbed from the
body cavity. The instant absorption of BPA into
the circulatory system is important when the
plasma half life is calculated, since compound
constantly entering the blood stream would influ-
ence the result. The plasma half life of BPA was
calculated to 3.75 h from a first-order exponential
decay regression model (r2=0.924). If the excre-
tion rates in the three compartments examined are
compared (Fig. 4) they can be ranked in the
following order: liver\plasma\muscle. The
fastest excretion rate is found in the liver where a
reduction in BPA of 92.3% occurred within the
first 12 h (relative to levels at 2 h). A somewhat
smaller excretion rate is observed in muscle, which
probably can be ascribed to the lack of metabolis-
ing enzymes. In the injection experiment, a lag
period between maximum plasma concentrations
of BPA and BPAGA is observed (Fig. 2B). This
delay in BPAGA production could, as mentioned
earlier, be ascribed to an increase in enzyme num-
ber or activity during the first 12 h.
In the present study, BPAGA quantification
was performed by comparison of unknown sam-
ples to
a synthesised standard. Since glu-
curonidated xenoestrogens are not commercially
available, quantification of the compounds has so
far relied on measurements of the parent com-
pounds following cleavage by the enzyme b-glu-
curonidase (Kolanczyk et al., 1999; Nakagawa
and Tayama, 2000). The direct measurement of
glucuronide is however easier to perform and
Fig. 5. Vitellogenin (VTG) concentrations in the blood plasma of male and female rainbow trout (mean9SEM). The fish were
either water-exposed to 0.44 mM BPA or intraperitoneally injection with 154 mmol BPA/kg. Group size varies between two and four
individuals. The solid and dashed lines indicate mean VTG concentrations of control male or female fish sampled throughout the
experiment. Asterisks indicate groups of fish significantly different from their respective control (PB0.05). The limit of quantifica-
tion was 40 ng VTG/ml.

C. Lindholst et al. / Aquatic Toxicology 55 (2001) 75–84
83
involves fewer analytical steps that may influence
the final result. Molecules containing two or more
hydroxyl groups, like BPA, could be doubly con-
jugated by UDP-glucuronosyl transferase. In rats,
the metabolism of brominated BPA results in 34%
diglucuronidated conjugates to be found in the
gall (Larsen et al., 2000). Diglucuronidated BPA
was however not measured in this experiment.
The present study shows a lag period of 3–5
days for females and 5–7 days for male fish
between BPA exposure through injection and a
significant increase in the VTG concentration
(Fig. 5B). During this lag period a possible gener-
ation of cellular components (e.g. the de novo
synthesis of additional oestrogen receptors) neces-
sary for the synthesis of vitellogenin may take
place (Mommsen and Walsh, 1988). Le Guellec et
al. (1988) were the first to describe this lag period
in male rainbow trout following acute and chronic
stimulation with estradiol. Though their results
were not subjected to any statistical treatment it
was evident that a several-fold induction in VTG
levels occurred after 7 days of chronic exposure.
An interesting aspect of the observed VTG re-
sponse is the finding that BPA is almost absent
from the liver at the time where maximal VTG
levels are found in the blood (120 and 168 h). This
result may suggest that the oestrogen receptors
located in the hepatocytes may remain active for a
certain time period after the release of the BPA
molecules. As a consequence of this continual
vitellogenesis, it is not possible to correlate actual
internal BPA concentrations with VTG levels in a
fluctuating environment. Furthermore, a sudden
rise in the estrogenic load of polluted waters may
cause an effect on the vitellogenesis several days
later. This persistent vitellogenic response could
reveal fluctuating estrogenic contents of polluted
waters not observable from chemical analysis of
water samples.
before a significant VTG response is observed
varies among males and females. In males, the
response is detected at 168 h whereas in females
significance can be seen at 120 h. Whether this
difference is caused by an enhanced ability of
juvenile females to induce vitellogenesis is uncer-
tain. However, this observation further demon-
strates that sex separation is necessary when the
vitellogenesis is used as a biomarker.
One of the aims of this study was to monitor
the metabolism of a xenoestrogenic compound.
As a representative of the metabolites generated
from BPA, the plasma level of BPAGA was mea-
sured throughout the experiment, showing con-
centration levels more than twice that of the
parent compound. This finding, together with the
fact that BPAGA is not the only degradation
product generated from BPA suggests that the
metabolism of xenoestrogens contributes to a sig-
nificant reduction of the overall estrogenic load of
the animal. The metabolism of xenoestrogens is
however influenced by factors such as their distri-
bution in the organism or availability in the circu-
latory system. Therefore further studies could
reveal the actual significance of metabolism possi-
bly explaining the relative potencies of various
endocrine disrupting compounds.
Acknowledgements
The authors wish to thank Birthe Christensen
for excellent technical assistance. Furthermore,
Jakob Bunkenborg from the Department of
Chemistry, Odense University, is thanked for his
assistance in the use of NMR. This work was
funded by a Danish Environmental Research Pro-
gramme grant to Professor Poul Bjerregaard and
Associate Professor Bodil Korsgaard.
In the present experiment, male and female fish
were divided into separate groups. This differenti-
ation was necessary due to the fact that VTG
levels differ between immature male and female
fish (Carlson and Williams, 1999). Apart from the
different control VTG levels, another sex differ-
ence is observed among fish from the injection
group (Fig. 5B). The duration of the lag period
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