The Impact on Reproduction of an Orally Administered Mixture of Selected PCBs in Zebrafish ( Danio rerio )

Abstract

Zebrafish (Danio rerio) were orally exposed to a mixture of 20 PCBs in three different dose levels (0.008, 0.08, and 0.4 microg of each congener per gram of freeze-dried chironomids). Generally, the PCBs accumulated in a dose-related manner. After 13 weeks of exposure body, liver, and ovary weights, as well as the liver and ovary somatic index, were significantly lower in exposed groups. In addition, the PCB mixture was an effective inducer of hepatic EROD activity. The reproduction study performed with exposed females and unexposed males after 9 weeks revealed that median survival time for larvae was only 7.7 days in the high-dose group as compared with 14 days in controls. Furthermore, egg production was reduced in all three groups exposed. No differences in hatching frequency or median hatching time were recorded. Histologically, females in both the intermediate and high-dose groups contained a reduced number of mature oocytes. The present study demonstrates that the potency of the mixture of selected PCBs induces hepatic EROD activity and has a clearly negative effect on zebrafish reproduction.

Full text

The Impact on Reproduction of an Orally Administered Mixture
of Selected PCBs in Zebrafish (Danio rerio)
S. O
¨
rn,
1
P. L. Andersson,
2
L. Fo¨rlin,
3
M. Tysklind,
2
L. Norrgren
1
1
Faculty of Veterinary Medicine, Department of Pathology, Swedish University ofAgricultural Sciences, S-750 07 Uppsala, Sweden
2
Umeå University, Institute of Environmental Chemistry, S-901 87 Umeå, Sweden
3
Department of Zoofysiology, University of Go¨teborg, S-413 90 Go¨teborg, Sweden
Received: 7 May 1997/Accepted: 29 November 1997
Abstract. Zebrafish (Danio rerio) were orally exposed to a
mixture of 20 PCBs in three different dose levels (0.008, 0.08,
and 0.4 µg of each congener per gram of freeze-dried chirono-
mids). Generally, the PCBs accumulated in a dose-related
manner. After 13 weeks of exposure body, liver, and ovary
weights, as well as the liver and ovary somatic index, were
significantly lower in exposed groups. In addition, the PCB
mixture was an effective inducer of hepatic EROD activity. The
reproduction study performed with exposed females and unex-
posed males after 9 weeks revealed that median survival time
for larvae was only 7.7 days in the high-dose group as
compared with 14 days in controls. Furthermore, egg produc-
tion was reduced in all three groups exposed. No differences in
hatching frequency or median hatching time were recorded.
Histologically, females in both the intermediate and high-dose
groups contained a reduced number of mature oocytes. The
present study demonstrates that the potency of the mixture of
selected PCBs induces hepatic EROD activity and has a clearly
negative effect on zebrafish reproduction.
Several studies indicate that PCBs (polychlorinated biphenyls)
affect reproductive success in wild fish populations. Lake trout
(Salvelinus namaycush) from the Great Lakes have been
reported to suffer from decreased hatching success and high
embryonic mortality (Mac et al. 1993). Hansen et al. (1985)
reported decreased viable hatch in Baltic herring (Clupea
harengus) and von Westernhagen et al. (1981) found impaired
egg development and survival in Baltic flounder (Platichtys
flesus). All of these reproductive failures have been associated
with high PCB levels in ovaries or eggs. Under laboratory
conditions, several studies have been conducted with commer-
cial PCB mixtures or single PCB congeners to evaluate the
effects on reproduction in fish. Bengtssson (1980) reported that
adult minnows (Phoxinus phoxinus) exposed to Clophen A50
suffered from delayed spawning and offspring showed reduced
hatching time and frequency. Later, Holm et al. (1993) reported
that three-spined stickleback (Gasterosteus aculeatus) exposed
to Clophen A50 suffered from reduced spawning success.
Furthermore, injection of a single PCB congener (3,38,4,48-
tetrachlorobiphenyl) in white perch (Morone americana) was
shown to impair both maturation of adult females and survival
of their offspring (Monosson et al. 1994). Other reproductive
anomalies observed upon PCB exposure include inhibition of
spermatogenesis and various testicular abnormalities (San-
galang et al.1981; Freeman et al. 1982) as well as disruption of
reproductive endocrine function (Khan and Thomas 1996).
However, only few long-term PCB exposure studies have
been reported and to the best of our knowledge there is none
performed on the zebrafish, which has become widely used in
ecotoxicological test systems. Furthermore, most studies con-
ducted to evaluate the effects of PCBs are, as described above,
based on exposure of single PCB congeners, e.g. #126 and
#169, or technical mixtures, such as ClophenA50 and Arochlor
1254. In order to cover the broad structural variation, tested
PCBs must be carefully preselected. The PCBs consist of a
large number of congeners, which represent a broad variation in
physicochemical characteristics (Andersson et al. 1996). By
using statistical design in combination with principal compo-
nent analysis, 20 PCBs have been selected to represent the 154
tetra- to hepta-chlorinated congeners (Tysklind et al. 1995). A
large spread in physicochemical characteristics of the included
PCBs determined the selection. The present study employs
these 20 PCBs, including congeners of each degree of chlorina-
tion, i.e., tetra to hepta, as well as each number of chlorine
atoms in ortho position.
The purpose of the present study was to evaluate reproduc-
tional effects in the zebrafish after long-term oral exposure to a
mixture of 20 selected PCB congeners.
Materials and Methods
Food Preparation
The mixture of selected PCBs contained equal amounts of each
congener and the purity, as controlled by HRGC/LRMS (MD800,
Fisons, UK), was more than 99%. The PCBs was purchased from Ultra
Scientific, North Kingstown, RI, USA (nos. 58, 78, 173, 188, and 190)
Correspondence to: L. Norrgren
Arch. Environ. Contam. Toxicol. 35, 52–57 (1998)
ARCHIVESOF
E
nvironmental
C
ontamination
and
T
oxicology
r
1998 Springer-Verlag New York Inc.

and fromAccuStandard, New Haven, CT, USA(nos. 41, 51, 60, 68, 91,
99, 104, 112, 115, 126, 143, 153, 169, 184, and 193). Numbering of the
PCBs throughout this paper is according to the IUPAC system
(Ballschmiter and Zell 1980; Schulte and Malisch 1983). The PCB
mixture was dissolved in 90 ml isooctane at different concentrations,
each of which was mixed with 30 g of freeze-dried chironomids
(Nutrafiny). The food was prepared in three dose levels of PCB, i.e.
low (Ld), intermediate (Id), and high (Hd). The control food was
treated with isooctane only. The solvent was evaporated at 50°C and 70
rpm using a rotary evaporator. The final concentrations in the low,
intermediate, and high dose levels were 0.008, 0.08, and 0.4 µg of each
congener/g food, respectively.
Experimental Design
Zebrafish, with a weight of 150–200 mg, were bought from a local pet
shop. After 4 weeks of acclimatization, the experiment started. The
exposure took place in 40-L aquariums, equipped with external filters
(EHEIM 2211). Once a month, one-fourth of the water volume and the
filter wadding was changed. The temperature was kept constant at
25°C, and the photoperiod was 10 h of light followed by 14 h of
darkness. The zebrafish were divided into four experimental groups,
each of 30 fish, i.e. one control group and one group for each dose level
of PCB. The fish were fed chironomids daily, in amounts equivalent to
about 2% of their body weight measured over the experimental period.
Sampling and Analysis
After 4 and 13 weeks of feeding, sampling took place. Fish were
randomly sampled to the sum of 10 females from each of the four
aquariums. At all samplings, body, liver, and gonadal weights were
recorded. For each female, the liver somatic index (LSI, liver
weight 3 100/total body weight) and gonad somatic index (GSI,
gonadal weight 3 100/total body weight) were calculated. Ovaries
were fixed in phosphate-buffered formalin, processed, and embedded
in paraffin. Serial sections were cut, stained with eosin-hematoxylin
and examined by light microscopy. In addition, female liver samples
were used for measurement of the cytochrome P450–dependent
ethoxyresorufin-O-deethylase (EROD) activity. The livers were
weighed, immediately frozen in liquid nitrogen, and kept at 270°C
until analysis. Homogenates were madefrom individual livers ultrasoni-
cated for5sin0.2mlof0.1Msodium phosphate buffer (pH 7.4)
containing 0.1 mM phenylmethylsulfonyl fluoride (PMSF), 0.02 mM
butylhydroxy toluene (BHT), 1 mM dithiothreitol (DTT), and 0.1 mM
EDTA. EROD activity was measured immediately after ultrasonication
as described by Andersson et al. (1985). Liver homogenate protein
content was measured according to Lowry et al. (1951), using bovine
serum albumin as standard.
Reproduction
After 9 weeks of exposure, the reproduction study was started. The
temperature was raised and kept constant at 27°C and the photoperiod
was 12:12 h of light:darkness. Rectangular spawning tanks were used
for the study and water was aerated using Rena 301 air pumps. Five
females from each experimental group were transferred to four
spawning tanks (one tank for each dose group), each tank containing
five net breeding traps, one per female, so that the females were
separated from each other. Each female was placed with two unexposed
males. Males were changed if spawning did not occur within a few
days. The traps were examined every day, at a specific time, and the
eggs were collected and transferred to 100-ml petri dishes. In cases
where there were large numbers of eggs, each batch was divided into
several groups so that each petri dish contained no more than 30 eggs.
Water was renewed every day in the dishes. The number of hatched and
surviving eggs/larvae were recorded daily in order to evaluate median
hatching time and median survival time. Parental fish were fed
unpolluted chironomids and daphnia during the reproduction experi-
ment whereas the larvae were not fed at all.
PCB Analysis
The female zebrafish from the two samplings, i.e. taken at 4 and 13
weeks, used for morphological studies and measurements of enzymatic
activity, were also analyzed for contents of PCBs. Since livers and
ovaries were dissected, the analysis of PCBs does not represent a
whole-body analysis. Each sample contained 10 females, which were
pooled at each dose level and time of exposure. The samples were
homogenized with 5–10 g of Na
2
SO
4
and transferred to a glass column
for extraction of lipids. Before extraction PCB no. 50 and 189 were
added as internal standards. Lipids were extracted with acetone:hexane
(5:2) followed by hexane:diethylether (9:1). After gravimetrical deter-
mination of lipids, a semipermeable membrane device (SPMD) was
used for dialysis of lipids using cyclohexane as solvent. Further
clean-up of the extracts included florisil-gel open chromatography and
an acid silica column. As a keeper tetradecane was added and the
extracts was evaporated to the final volume. Before analysis, PCB no.
199 was added as recovery standard.Analysis of the PCBs were carried
out by HRGC/ECD (Hewlett Packard 6890). For identification and
quantification of the PCBs in the mixture a standard was run. The
tetra-chlorinated biphenyls were quantified versus PCB no. 52, the
penta CBs vs. no. 101, the hexa CBs vs. no. 153, and the hepta CBs vs.
no. 180.
Statistical Analysis
Statistical analyses of LSI, GSI, EROD, body weight, liver weight, and
ovary weight were conducted using the nonparametrical Mann-
Whitney U test, comparing the exposed groups to the control group.
The significance level was set at 0.95 (p # 0.05). The symbols *, **,
and *** represent p values of p # 0.05, p # 0.01, and p # 0.001,
respectively. In the reproduction study, median hatching time and
median survival time were calculated using a simple regression plot
(Statview 4.0 for Macintosh) and tested for statistical significance
using the Mann-Whitney U test, as described above.
Results and Discussion
In the present study, zebrafish were orally administered a
mixture of selected PCBs for a total period of 13 weeks. Only
few experimental long-term exposure studies are reported in the
literature concerning effects of PCBs on fish. In the natural
environment, fish are continuously exposed to a broad range of
toxicants over their whole lifetime. Therefore it is important to
focus on toxicological effects observed during long times of
exposure. There are several reports concerning PCB residues in
ovaries and the effects on reproduction in feral fish. Baltic
flounder (Platichtys flesus) eggs with PCB concentration exceed-
ing 0.12 µg/g (wet wt) showed a reduction in viable hatch (von
Westernhagen et al. 1981). Furthermore, when the concentra-
tion was near or higher than 0.25 µg/g, viable hatch was lower
than 15%. In whiting (Merlangius merlangus), von Westernha-
gen et al. (1989) reported the critical ovarian threshold level for
PCB to be 0.2 µg/g. The PCBs in the present study consisted of
53PCBs and D. rerio Reproduction

20 structurally diverse tetra- to hepta-chlorinated congeners.
The chemical analysis of the PCBs, with recoveries of 68–90%,
was conducted for pooled females sampled after 4 and 13
weeks, from the different dose groups. Generally, the concentra-
tions of each congener increased by the dose given and with
time of exposure. The levels found of each congener vary
greatly depending on structure-specific characteristics. Conge-
ners lacking chlorine atoms in vicinal meta and para position,
such as PCB no. 104 and 143, as well as the non-ortho
substituted PCBs, e.g. no. 126 and 169, are found in lower
concentrations. These phenomena have recently been reported
for three-spined sticklebacks (van Bavel et al. 1996). After 4
and 13 weeks of exposure, SPCB concentrations for control,
Ld, Id, and Hd groups were 0.06, 0.11, 0.58, and 1.9, and 0.07,
0.14, 1.1, and 2.7 µg/g fish (wet wt), respectively. The tissue
concentration of PCBs in the Id group increased twofold over
the 4- to 13-week exposure period, whereas the Hd group
increased only 1.4-fold over the same period. The lower
increase of the Hd group might indicate a concentration near
equilibrium of accumulation and clearance rate.
Monosson et al. (1994) reported that female white perch
(Morone americana) accumulated PCB no. 77 in ovary .
liver . skeletal muscle after a single intraperitoneal (IP)
injection. Furthermore, the residues of PCB#77 were 10–15-
fold higher in the ovaries and three–five-fold higher in livers, as
compared with skeletal muscle. In a study similar to the present
(Holm et al. 1993), sticklebacks were exposed orally to two
different dose levels of Clophen A50. After 3.5 months of
exposure, the whole-body PCB residues of the female stickle-
backs were reported to be 102 and 289 µg/g (wet wt),
respectively. These concentrations were 50–100-fold higher
than those of the Hd group (13 weeks) in the present zebrafish
study (2.7 µg/g wet wt). However, in the present study the
actual body burden was probably much higher than that
recorded since livers and ovaries were dissected prior to
analysis. Probably, these tissues contained higher levels of
PCBs than that recorded in body concentration. In comparison,
in minnows (Phoxinus phoxinus) exposed to Clophen A50,
Bengtsson (1980) reported abnormally high egg mortality at
ovary residue levels of 15 µg/g. Furthermore, Nebeker et al.
(1974) reported a reduction in reproductive success when
fathead minnow (Pimephales promelas) and flagfish (Jor-
danella floridae) reached a total body burden of 92 µg/g PCB.
Measurement of EROD (CYP1A) activity is a commonly
used assay in the assessment of exposure of organisms to
persistent organic pollutants, such as PCDDs, PCDFs, PCNs,
and PCBs (Goksøyr and Fo¨rlin 1992). The present study shows
that the PCB mixture used is an effective inducer of the hepatic
cytochrome P450 system in zebrafish (Table 1). In a study by
Buchmann et al. (1993), where zebrafish were exposed to
TCDD (2,3,7,8-tetrachlorodibenzo-p-dioxin), the unexposed
controls showed an EROD activity of 9.5 pmol · mg
protein
21
· min
21
. This is in accordance with the controls in the
present study, where the first and second sampling showed 14.4
and 5.2 pmol · mg protein
21
· min
21
in mean hepatic EROD
activity, respectively. After 4 weeks of exposure, a 12-fold
elevated EROD activity was recorded in the Hd group com-
pared with controls. The second sampling revealed a significant
dose-dependent EROD induction (Table 1). However, in the Hd
group a reductionin EROD wasrecorded, probably dueto some
adaptory response or intracellular damage of the hepatocytes.
Table 1. HepaticEROD activity (pmol resorufin · min
21
· mg protein
21
)
in female zebrafish after 4 and 13 weeks of PCB exposure
a
Dose Group
4 Weeks of PCB
Exposure
13 Weeks of PCB
Exposure
Control 14.4 6 10 (n 5 10) 5.20 6 1.8 (n 5 10)
Low 6.50 6 3.4* (n 5 9) 9.26 6 4.1* (n 5 8)
Intermediate 18.2 6 16 (n 5 9) 22.0 6 11*** (n 5 10)
High 170 6 80*** (n 5 10) 49.1 6 61** (n 5 6)
a
Values are given as mean 6 standard deviation (SD)
Table 2. Total mortality (n 5 30) during the experiment and body
weights recorded after 4 and 13 weeks of PCB exposure
a
Dose
Group
(n 5 10)
Mor-
tality
(%)
4 Weeks of
PCB Exposure
13 Weeks of
PCB Exposure
Body
Weight
(mg)
Weight
Index
(%)
Body
Weight
(mg)
Weight
Index
(%)
Control 3 235 6 47 127 402 6 110 1117 (171)
Low 3 253 6 59 137 299 6 70* 162 (118)
Inter-
mediate 10 279 6 87 151 265 6 46** 143 (25)
High 20 263 6 98 142 224 6 71** 121 (215)
a
Weight indices are related to fish weight before exposure (185 6 36
mg) and indices in parenthesis between samplings. Values are given as
mean 6 standard deviation (SD)
Table 3. Ovary weights and gonad somatic indices (GSI) in female
zebrafish following 4 and 13 weeks of PCB exposure
a
Dose
Group
(n 5 10)
4 Weeks of
PCB Exposure
13 Weeks of
PCB Exposure
Ovary
Weight
(mg)
GSI
(%)
Ovary
Weight
(mg)
GSI
(%)
Control 16.6 6 11 6.84 6 4.0 37.7 6 25 8.90 6 4.8
Low 12.4 6 11 5.02 6 4.8 21.1 6 18 6.47 6 5.6
Inter-
mediate 19.4 6 25 5.89 6 5.7 7.51 6 7.0** 2.92 6 2.8**
High 23.6 6 25 7.67 6 5.9 3.62 6 4.9*** 1.36 6 1.4***
a
Values are given as mean 6 standard deviation (SD)
Table 4. Liver weights and liver somatic indices (LSI) in female
zebrafish after 4 and 13 weeks of PCB exposure
a
Dose
Group
(n 5 10)
4 Weeks of
PCB Exposure
13 Weeks of
PCB Exposure
Liver
Weight
(mg)
LSI
(%)
Liver
Weight
(mg)
LSI
(%)
Control 4.3 6 2.1 1.84 6 0.81 11 6 4.5 2.74 6 0.61
Low 4.8 6 2.2 1.83 6 0.56 6.3 6 2.9* 2.00 6 0.66*
Inter-
mediate 5.2 6 2.6 1.79 6 0.67 2.9 6 1.5*** 1.13 6 0.69***
High 5.1 6 3.1 1.91 6 0.91 2.8 6 1.8*** 1.17 6 0.46***
a
Values are given as mean 6 standard deviation (SD)
54 S. O
¨
rn et al.

Inhibition of CYP1A1 by PCB congeners of commercial
mixtures have recently been reported. In scup liver (Stenotomus
chrysops), PCB no. 77 has been indicated to simultaneously
cause catalytic inhibition, induction of CYP1A1 mRNA, and
inhibition of CYP1A1 protein (White et al. 1997). In rainbow
trout, IP injection with Clophen A50 altered the responsiveness
toward PCB, measured as a reduction in CYP1A1 mRNA and
CYP1A1 protein (Celander and Fo¨rlin 1995). The reason for
this inhibition is not known, but it is suggested in both reports
that it can be an effect directly on the CYP1A1 system by PCB.
During the exposure period, a low mortality in the control
and Ld groups, with one fish (3%) dead in each group, was
recorded. The numbers of deaths in the Id and Hd groups were
three (10%) and six fish (20%), respectively (Table 2). No
significant differences were recorded after 4 weeks of PCB
exposure in body, liver, or ovary weights between controls and
exposed groups. However, after 13 weeks of exposure, the
mean body weight was almost half in the Hd group, 224 mg as
compared with 402 mg in control (Table 2). The decreases in
body weights in exposed groups are mostly due to decreases in
GSIs (Table 3). Furthermore, there was also decreases in liver
weights and LSIs in the Id and Hd groups (Table 4), which may
be due to metabolic stress and toxic effects, leading to
degeneration of the liver. In addition, ovary weights and GSIs
were found to be significantly lower in the Id and Hd groups
(Table 3). The GSIs in females in the Id and Hd groups were
approximately one-third and one-sixth of that in the controls,
respectively. This is in accordance with a study by Monosson et
al. (1994), where adult white perch were given IP injections of
PCB no. 77 prior to the spawning season. The results showed
that fewer females matured in the group receiving the highest
dose (3 IP injections of 5 µg/kg) and those that did mature had a
GSI approximately half that of control females. The study also
showed that females from all groups were able to spawn, but
there was a reduction in egg deposition in the groups exposed to
PCB no. 77. This is in accordance with observations of
reproduction disturbances within the present study. These
results suggest that the PCB mixture, or certain PCB congeners,
has a suppressive effect on oocyte maturation rather than
affecting ovarian development. This is also supported by the
Fig. 1. Ovaries from (a) control ze-
brafish with oocytes in different stages
of maturity and (b) from the high dose
group containing only primary oocytes
(3 110)
Table 5. Spawning success, median hatching time, and median survival timein the reproduction study, which includedunexposed males and female
zebrafish exposed to three different dose levels of PCB for 9 weeks
a
Dose
Group
Spawning
Median Hatching
Time (days)
Median Survival
Time (days)
Spawning
Females
(%)
Mean #
Eggs per
Female
Early
Mortality
(%)
Control 80 238 0 2.98 6 0.42 (n 5 255) 13.8 6 1.2 (n 5 178)
Low 40 220 12 2.81 6 0.57 (n 5 114) 13.3 6 1.2 (n 5 98)
Intermediate 60 167 9 2.84 6 0.63 (n 5 70) 13.1 6 1.9 (n 5 19)
High 60 96 5 2.63 6 0.43 (n 5 123) 7.67 6 0.49*** (n 5 92)
a
Values are given as mean 6 standard deviation (SD)
55PCBs and D. rerio Reproduction

histological examination, which showed that after 13 weeks of
exposure, both Idand Hd groupscontained a reducednumber of
mature oocytes compared with the control group (Figure 1).
Selman et al. (1993) divided oocyte development in zebrafish
into five different stages (I–V). According to this classification,
the oocytes, especially in the Hd group, seemed to be arrested in
stage I (Figure1), containing nocortical alveoli, noyolk bodies,
and with few nucleoli in the nucleus. The exact mechanism by
which PCBs negatively affect the reproduction system is
unknown. Atlantic croaker (Micropogonias undulatus), fed the
technical PCB mixture Aroclor 1254, were reported to have
decreased plasma estradiol and vitellogenin concentrations, as
well as decreased GSI (Thomas 1989). Furthermore, Thomas
(1989) also observed a decreased secretion of gonadotropin
from the pituitary when incubated in vitro after the in vivo
exposure. PCBs might have a site in the hypothalamus-pituitary
complex, and thereby decrease the secretion of gonadotropins
(Thomas 1989). This would lead to decreased estradiol produc-
tion and plasma vitellogenin concentration, which reduce
oocyte maturation.Another factor that affect reproduction is the
stress hormone cortisol. Cortisol has been shown to inhibit
testosterone and estradiol production in rainbow trout ovarian
follicles (Carragher and Sumpter 1990). Elevated plasma
cortisol levels and EROD activity have recently been reported
in rainbow trout exposed to PCB no. 77 (Vijayan et al. 1997).
Therefore, the toxicological stress due to PCB exposure might
have a suppressive effect on oocyte maturation.
When considering the spawning, there was a dose-related
reduction of the number of eggs deposited per female (Table 5).
This is in accordance with the histological evaluation showing a
reduction of mature oocytes in the exposed groups. The control
group showed the largest number of spawning females (80%),
producing a mean of 238 eggs per female. The least number of
eggs deposited was from the females in the Hd group, less than
half of that in controls, followed by the Id and Ld groups,
respectively (Table 5). The dose-related reduction of eggs laid
in the exposed groups is in accordance with previous studies.
Wannemacher et al. (1992) reported that, in female zebrafish,
exposure to 5, 10, and 20 ng TCDD/fish led to a dose-related
reduction of the number of eggs. Similar results were obtained
in a study by Bengtsson (1980) where Clophen A50 was shown
to reduce the number of spawning occasions in minnows. In the
present study, the numbers of deaths occurring within the first
three days (early mortality), were recorded to be 0% in controls,
12% in the Ld group, and 9 and 5% in the Id and Hd groups,
respectively (Table 5). We have no explanation for the fact that
the low-dose group produced more vulnerable embryos; how-
ever, it could be due to individual female variation in egg
quality. No differences in hatching frequency were observed,
with an almost 100% hatching success in all groups. Further, no
statistical difference could be found in median hatching time
(Table 5). However, the embryos of the Hd group tended to
hatch earlier than embryos of the other groups. Although not
statistically significant (p 5 0.14), median hatching time of the
Hd group was recorded as 2.63 days compared with 2.98 days
in the control group.
When considering median survival time of the larvae, there
was a dramatic decline in the Hd group. Larvae of the control
group showed a median survival time of 13.8 days, compared
with only 7.67 days (p 5 0.0007) in the Hd group (Table 5).
Maternal exposure to PCB is well known to be associated with
decreased larval survival (von Westernhagen et al. 1981; Black
et al. 1988; Monosson et al. 1994). In the study by Monosson et
al. (1994), a decline in survival near the end of the yolk-sac
absorption was reported in larvae from female white perch
exposed IP to 1.0 and 5.0 µg/g of PCB no. 77. Walker et al.
(1992) exposed eggs from rainbow trout (Oncorhynchus mykiss)
and lake trout (Salvelinus namaycush) to TCDD, both by an
egg-injection method and waterborne exposure, andthe predomi-
nantly overall mortality occurred during the yolk-sac stage, not
during the egg stage.
In conclusion, the selected PCBs accumulated in a dose-
related manner. It is demonstrated that the levels of PCBs used,
when exposed to zebrafish, affects reproduction negatively,
with disturbances in oocyte maturation and offspring survival.
Furthermore, the mixture of PCBs strongly induces hepatic
EROD activity.
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