Aquatic Toxicology 106–107 (2012) 173–181
Contents lists available at SciVerse ScienceDirect
Aquatic Toxicology
journal homepage: www.elsevier.com/locate/aquatox
Disruption of endocrine function in in vitro H295R cell-based and in in vivo assay
in zebrafish by 2,4-dichlorophenol
Yanbo Ma a,b , Jian Han a , Yongyong Guo a , Paul K.S. Lam c , Rudolf S.S. Wu d , John P. Giesy c,d,e,f,g ,
Xiaowei Zhang h , Bingsheng Zhou a,∗
a
State Key Laboratory of Freshwater Ecology and Biotechnology, Institute of Hydrobiology, Chinese Academy of Sciences, Wuhan 430072, China
Graduate School of the Chinese Academy of Sciences, Beijing 100039, China
c
Department of Biology and Chemistry, City University of Hong Kong, Kowloon, Hong Kong SAR, Hong Kong, China
d
School of Biological Sciences, the University of Hong Kong, Hong Kong SAR, Hong Kong, China
e
Department of Veterinary, Biomedical Sciences, University of Saskatchewan, Saskatoon, Canada
f
Department of Veterinary Biomedical Sciences and Toxicology Centre, University of Saskatchewan, Saskatoon, Canada
g
Zoology Department, College of Science, King Saud University, P.O. Box 2455, Riyadh 11451, Saudi Arabia
h
State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing, China
b
a r t i c l e
i n f o
Article history:
Received 22 August 2011
Received in revised form
12 November 2011
Accepted 15 November 2011
Keywords:
Hormones
In vitro
In vivo
H295R
HPG axis
Zebrafish
a b s t r a c t
Chlorophenols in the aquatic environment have been of concern due to their potential effects on human
and wildlife. In the present study, the endocrine disrupting effects of 2,4-dichlorophenol (2,4-DCP) were
investigated in vitro and in vivo. In the in vitro assay, H295R human adrenocortical carcinoma cells were
used to determine the potential effects of 2,4-DCP on steroidogenesis. Exposure to 0, 0.1, 0.3 or 1.0 mg
2,4-DCP/L resulted in less production of 17␤-estradiol (E2) and alterations in transcript expressions of
genes involved in steroidogenesis, including cytochrome P450 (CYP11A, CYP17, CYP19), 3ˇHSD, 17ˇHSD
and StAR. In the in vivo study, effects of 0, 0.03, 0.1 or 0.3 mg 2,4-DCP/L on concentrations of steroid
hormones in plasma of adult zebrafish (Danio rerio) were measured and expression of mRNA of selected
genes in hypothalamic-pituitary-gonadal (HPG) axis and liver were determined. Exposure of zebrafish
to 2,4-DCP resulted in lesser concentrations of E2 accompanied by down-regulation of CYP19A mRNA
in the females. In males, exposure to 2,4-DCP resulted in greater concentrations of testosterone (T) and
E2 along with greater mRNA expression of CYP17 and CYP19A. The mRNA expression of prostaglandin
synthase (Ptgs2) gene, which regulates ovulation, was down-regulated in females, but up-regulated in
males. The hepatic estrogenic receptor (ER˛ and ERˇ) and vitellogenin (VTG1 and VTG3) mRNAs were upregulated in both females and males. The average number of eggs spawned was significantly less upon
exposure to 2,4-DCP. Exposure of adult zebrafish to 2,4-DCP resulted in lesser rates of hatching of eggs.
The results demonstrated that 2,4-DCP modulates transcription of steroidogenetic genes in both H295R
cells and in the zebrafish HPG-axis and disrupts steroidogenesis, which in turn, can cause adverse effects
on reproduction in fish.
© 2011 Elsevier B.V. All rights reserved.
1. Introduction
Extensive use of chlorophenols (CPs) as a biocide, wood treatment agent, and as a byproduct of bleaching in paper mills
have resulted in CPs being distributed in the global environmental (Stringer and Johnston, 2001). Because CPs can cause
adverse effects on human and wildlife, including chronic toxicity,
mutagenicity and carcinogenicity, the United States Environmental Protection Agency (US EPA) has classified pentachlorophenol
(PCP), 2,4,6-trichlorophenol (2,4,6-TCP), and 2,4-dichlorophenol
∗ Corresponding author. Tel.: +86 27 68780042; fax: +86 27 68780123.
E-mail address: bszhou@ihb.ac.cn (B. Zhou).
0166-445X/$ – see front matter © 2011 Elsevier B.V. All rights reserved.
doi:10.1016/j.aquatox.2011.11.006
(2,4-DCP) as priority pollutants (Ramamoorthy and Ramamoorthy,
1997). Of these chemicals, 2,4-DCP is the most abundant in
aquatic environments (House et al., 1997). 2,4-DCP is primarily
formed as a biotransformation product of the pesticide 2,4dichlorophenoxyacetic acid (Zona et al., 2002) and is also derived as
a degradation product of more chlorinated CPs (Brillas et al., 2000).
Widespread occurrences of 2,4-DCP in surface waters have been
reported in several countries at concentrations ranging from <1
to 4.7 ␮g/L (House et al., 1997; Chiron et al., 2007). In China, concentrations of 2,4-DCP as great as 20.0 ␮g/L have been observed in
surface waters of seven major watersheds in three drainage areas
(Gao et al., 2008).
Previous in vitro and in vivo studies have found that 2,4-DCP
can modulate the endocrine system. For example, by use of a yeast
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Y. Ma et al. / Aquatic Toxicology 106–107 (2012) 173–181
two-hybrid transactivation reporter gene assay, 2,4-DCP was
observed to cause effects mediated through the ERE (estrogen responsive element) in a concentration-dependent manner
(Nishihara et al., 2000). In another reporter gene assay, 2,4-DCP
antagonized the androgen receptor (AR) (Li et al., 2010). 2,4-DCP
has also been shown to affect expression of estrogenic receptors
(ERs) and induction of vitellogenin (VTG) in an in vivo study of
the rare minnow (Gobiocypris rarus) (Zhang et al., 2008). However,
the potential mode of action for reproductive toxicity of 2,4-DCP
is unknown and there is still a lack of understanding on whether
these molecular responses can be manifested to impairment of
reproduction.
Steroid hormones regulate reproductive processes. They have
direct effects on gametogenesis and reproductive maturations.
Endocrine disruptors can directly interact with receptors or alter
enzymes involved in steroid hormone synthesis and metabolism,
and thus, impair reproduction. Chemicals can alter expression of
steroidogenic genes or enzyme activities and thereby have the
potential to alter concentrations of hormones in blood and tissues (Hilscherova et al., 2004). In this regard, the steroidogenesis
assay based on H295R human adrenocortical carcinoma cells has
been developed for quantitative evaluation of xenobiotic effects on
transcription of genes involved in steroidogenesis (Sanderson et al.,
2000; Gracia et al., 2007; Harvey et al., 2007). The H295R assay has
been validated by the US EPA for use in a tiered screening approach
(Gracia et al., 2006). This system assay has the potential as a screening tool to discern the mechanisms of action of specific endocrine
modulating compounds. Modulation of transcript expression of
steroidogenic genes by other toxicants, such as dioxins and polychlorinated biphenyls (PCBs) (Li and Wang, 2005), bromophenols
(Ding et al., 2007), polybrominated diphenyl ethers (PBDEs) (He
et al., 2008), fungicides (Ohlsson et al., 2009), and various model
chemicals (Zhang et al., 2005) have been investigated. It has been
previously shown that pentachlorophenol or 2,4,6-trichlorophenol
could disrupt steroidogenesis in H295R cells via a cAMP-dependent
pathway (Ma et al., 2011). These results suggested that examination
of transcript expression of genes could be used to rapidly screen
for the potential disruption of steroidogenesis by various types of
toxicants.
Reproduction in fish is primarily regulated by the hypothalamicpituitary-gonadal axis (HPG) and steroidogenesis of gonad tissues
plays an important role in control of reproductive processes
(Richards, 1994; Nagahama and Yamashita, 2008; Sofikitis et al.,
2008). Theoretically, disruption at any point in this axis can
adversely affect the function of endocrine system and affect reproduction. Experiments with model chemicals, such as fadrozole,
prochloraz and ketoconazole, have been conducted to test the
impact on the HPG axis in fathead minnow (Pimephales promelas) (Villeneuve et al., 2007a) and Japanese medaka (Oryzias latipes)
(Zhang et al., 2008a,b).
In vitro assays have been regarded as simple, rapid and costeffective methods over in vivo techniques for assessing toxicity
of xenobiotics in animals. However, a challenge is to validate
results from in vitro assays on their relevance, sensitivity and predictability of in vivo assays (Segner et al., 2003). Therefore, the
objective of the present study was (1) to evaluate the effect of
2,4-DCP on steroidogenesis in the H295R cell; (2) for comparison purposes, we also tested in vivo assay to evaluate the impact
of 2,4-DCP on fish. In the in vitro assay, production of the steroid
hormones testosterone (T) and 17␤-estradiol (E2), as well as transcript expression of genes encoding for key steroidogenic enzymes
in the steroidogenic pathway (StAR, CYP11A, 3ˇHSD, CYP17, CYP19,
17ˇHSD) was examined. In the in vivo assay, effects of 2,4-DCP
on mRNA expressions of genes in HPG axis, steroid hormone
levels and reproduction were also investigated using zebrafish
(Danio rerio).
2. Materials and methods
2.1. Chemicals
2,4-Dichlorophenol (98.9%, CAS No. 120-83-2) was purchased
from AccuStandard Inc. (New Haven, CT, USA). 2,4-DCP was dissolved in dimethyl sulphoxide (DMSO) to form a stock solution
(100,000 mg/L) and stored at 4 ◦ C. All other chemicals used in this
study were of analytical grade.
2.2. H295R cell culture and chemical exposure
H295R cells were maintained in DMEM/F12 medium containing 1% insulin-transferring sodium selenite plus Premix (ITS) (BD
Bioscience, Bedford, USA), 2.5% Nu-Serum (BD Bioscience, Bedford,
USA), 100 U/mL penicillin, and 100 ␮g/mL streptomycin. The cells
were grown at 37 ◦ C with 5% CO2 . For determining the effects of
2,4-DCP on mRNA expression of genes and production of hormones,
cells were grown in 12-well plates, and 2 mL of cell suspension with
a density of 3 × 105 cells/mL was added to each well. After 24 h,
cells were exposed to 0, 0.1, 0.3 or 1.0 mg 2,4-DCP/L for 48 h. Three
replicates of each exposure were used in each experiment. The
exposure and control groups all received 0.1% (vol/vol) DMSO. After
48 h exposure, the culture medium was transferred to an Eppendorf
tube and stored at −80 ◦ C for quantification of hormones.
For the cell viability assay, the cells were gown in 24-well
culture plate with a density of 3 × 105 cells/mL. Cells were examined after 48 h by measuring LDH leakage as previously described
(Ma et al., 2011) utilizing a commercial kit (GMS 10073, GenMed
Scientifics Inc.). Three replicates were used in each treatment. No
significant LDH leakage was found in any of the 2,4-DCP treatments
(data not shown), suggested that there is no significant effect on
the cell viability.
2.3. Zebrafish maintenance and chemical exposure
Adult 3-month-old, wild type zebrafish (AB strain) were maintained in 15-L glass tanks, which contained 10 L charcoal-filtered
tap water, pH 7.0–7.4 at 28 ± 0.5 ◦ C with a 14:10 light/dark cycle
according to the previous method (Liu et al., 2009). Before exposure to 2,4-DCP, zebrafish were acclimated in tanks for one week
and each tank contained five females and five males. Zebrafish
were exposed to 0, 0.03, 0.1 or 0.3 mg 2,4-DCP/L for three weeks in
semi-static systems. Both the control and exposure groups received
0.027% (vol/vol) DMSO. Fish were fed Artemia nauplii twice daily.
Half of the exposure water in each tank was replaced daily with
fresh solution at the appropriate concentration. There were three
replicate tanks for the control and exposure groups.
2.4. Gamete parameters
Gamete quality is an important factor contributing to successful production of fish larvae. Both decreased gamete quality and
quantity will impair fertilization success, embryo development and
larval viability. In the last two weeks of exposure, sexually mature
male and female zebrafish were paired and eggs were collected
according to the method described by Liu et al. (2010a). Eggs that
had been collected from the mating pairs were counted. Fecundity
was reported as the cumulative average number of eggs produced
per female per day (Ankley et al., 2005). On the last day, diameter and total protein content were determined from a sample of 15
eggs from each tank. The diameter of eggs was evaluated using an
Olympus IX71 microscope (Olympus America, Melville, NY, USA)
with a digital camera, and the image was examined by the use of
Image-Pro Plus 6.0 software (Liu et al., 2010a). The concentration
of protein in eggs was determined by the Bradford method using
bovine serum albumin (Sigma, St. Louis, MO, USA) as a standard.
Y. Ma et al. / Aquatic Toxicology 106–107 (2012) 173–181
175
Table 1
Primer sequences for the genes tested in the present study in H295R and zebrafish.
Gene name
Sequence of the primer (5 –3 )
Genbank accession no
Forward
Reverse
H295R
ˇ-Actin
CYP11A
StAR
3ˇHSD
CYP17
CYP19
17ˇHSD
caccttccagccttccttcc
gagatggcacgcaacctgaag
gtcccaccctgcctctgaag
tgccagtcttcatctacaccag
agccgcacaccaactatcag
aggtgctattggtcatctgctc
tgcgggatcacggatgactc
aggtctttgcggatgtccac
cttagtgtctccttgatgctggc
catactctaaacacgaaccccacc
ttcccagaggctcttcttcgtg
tcaccgatgctggagtcaac
tggtggaatcgggtctttatgg
gccaccattctcctcacaactc
NM
NM
NM
NM
NM
NM
NM
Zebrafish
rpl8
CYP19B
FSH
CYP17
CYP19A
3ˇHSD
17ˇHSD
ER˛
ERˇ
FSHR
ptgs2
VTG1
VTG3
ttgttggtgttgttgctggt
ggcagtctctggaggatgac
tgagcgcagaatcagaatg
gacagtcctccgcacatct
ctgaaagggctcaggacaa
agtgtcgcacatcgtctcag
aatagagggcgcttgtgaga
tgagcaacaaaggaatggag
tgattagctgggcgaaga
tccactcgctctttcgt
tggatctttcctgggtgaagg
tccattgctgaaaacgacaa
ggtggttcttggacttggtt
ggatgctcaacagggttcat
cagtgttctcgaagttctcca
aggctgtggtgtcgattgt
gcatgatggtggttgttca
tggtcgatggtgtctgatg
cagtcggaccagcttttctc
tccagctccttgtctccagt
gtgggtgtagatggagggttt
tatccagccagcagcatt
gttctcggacaccactattc
gaagctcaggggtagtgcag
tgcattcagcacacctctca
cacaggagaggatgggattt
NM 200713
AY780257
AY424303
AY281362
AF226620
AY279108
NM 205584
NM 152959
AJ414566
AY424301
AY028585
AF406784
AF254638
2.5. Rate of hatching
On the last day of exposure, one hundred randomly selected
fertilized eggs from each tank were collected and separately placed
in glass dishes (500 mL), containing 250 mL fresh water without
chemical exposure until 4 days post-fertilization (dpf) under static
conditions, and hatching rate was determined. Half of the water
was renewed daily.
2.6. Quantification of hormones
Extraction and quantification of hormones in the culture
medium of H295R cells or plasma of zebrafish, were performed
as previously described (Ma et al., 2011). Briefly, 0.5 mL culture
medium was extracted twice with diethyl ether (2.5 mL) and
the solvent was evaporated under a gentle stream of nitrogen.
The residue was dissolved in 250 ␮L ELISA buffer and hormones
in the culture medium were determined by competitive ELISA
according to the manufacturer’s instructions (Cayman Chemical
Company, Ann Arbor, MI; Testosterone [Cat # 582701], 17␤estradiol [Cat # 582251]). The intra-assay variation was determined
by including multiple microplate wells per sample and interassay variability was estimated by including the same samples
on multiple plates. Coefficients of variation (CV) were used to
describe the reproducibility of the results (The CV = (standard deviation/mean) × 100%). Intra-assay and inter-assay coefficients of
variation were <10% for testosterone and 17␤-estradiol. The detection limits was 6 pg/mL for T and 19 pg/mL for E2.
After exposure for 21 d, the zebrafish were anesthetized with
0.03% tricaine methyl sulphonate (MS-222) and masses determined. Blood was collected from the caudal vein of each fish, eight
individual fish from the three tanks were randomly sampled and
blood from 2 individual fish of the same sex was pooled to form a
composite sample. The pooled blood was centrifuged at 7000 × g
for 5 min at 4 ◦ C, and the supernatant was collected and stored at
−80 ◦ C. Before determining the concentrations of hormones, the
plasma collected was extracted twice with diethyl ether. Briefly,
the supernatant sample (female10 ␮L; male 7.5 ␮L) were diluted to
400 ␮L with UltraPure water and extracted twice with 2 mL diethyl
ether at 2000 × g for 10 min. The solvent used to extract hormones
was evaporated under a stream of nitrogen, and the residue was dissolved in 120 ␮L ELISA buffer. Hormones in plasma were measured
001101
000781
000349
000198
000102
000103
000414
by competitive ELISA by use of the methods suggested by the manufacturer (Cayman Chemical Company, Ann Arbor, MI, USA).
2.7. Quantitative real-time PCR assay of mRNA
Extraction of RNA from H295R cells, determination of purity of
the RNA, synthesis of first-strand cDNA and quantitative real-time
PCR were performed by use of previously described methods (Ma
et al., 2011). In zebrafish, after blood was collected, the gonad, liver
and brain were dissected and preserved in TRIzol reagent (Invitrogen, Carlsbad, CA, USA) for RNA sample preparation, first-strand
cDNA synthesis, and quantitative real-time polymerase chain reaction (q-RT-PCR) assay according to previously described protocols
(Liu et al., 2009). The q-RT-PCR was performed by use of the
SYBR Green PCR kit (Toyobo, Tokyo, Japan) on ABI 7300 System (PerkinElmer Applied Biosystems, Foster City, CA, USA). The
primer sequences of the selected genes were obtained by using the
online Primer 3 program (http://frodo.wi.mit.edu/) and are shown
(Table 1). The mRNA expression of each target gene was normalized to its corresponding rpl8 (ribosomal protein L8) mRNA content
(Uren-Webster et al., 2010). The rpl8 was selected as an internal
standard because in previous studies mRNA expression of this gene
in the liver and gonad tissues in fish was unaffected by exposure to
chemicals (Filby and Tyler, 2007). Using geNorm analysis, our studies also demonstrated that rpl8 is the most stable gene among the
five commonly used housekeeping genes (rpl8, 18S rRNA, ˇ-actin,
elongation factor 1␣, ef1␣, glyceraldehyde-3-phosphate dehydrogenase, gapdh) in both male and female gonad and the mRNA
expression of the gene is not affected after 2,4-DCP exposure (data
not shown). The relative mRNA expression was determined by the
2−CT method (Livak and Schmittgen, 2001).
2.8. Statistical analysis
The data was checked for normality by use of the
Kolmogorov–Smirnov test, and if necessary, data were logtransformed to approximate normality. Homogeneity of variances
was analyzed by Levene’s test. Differences between and among
treatments were evaluated by use of a one-way analysis of variance
(ANOVA) test followed by a Tukey’s multiple range tests using
SPSS 13.0 (SPSS, Chicago, IL, USA). The criterion for statistical
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Y. Ma et al. / Aquatic Toxicology 106–107 (2012) 173–181
Table 2
Somatic indices in zebrafish after exposure to 2,4-DCP (0, 0.03, 0.1, 0.3 mg/L) for 21 days.
Female
Weight (g)
HSIa
GSIb
Male
0
0.03
0.1
0.3
0
0.03
0.1
0.3
0.59 ± 0.06
3.19 ± 0.66
17.86 ± 0.78
0.48 ± 0.06
4.45 ± 0.89
16.19 ± 2.42
0.54 ± 0.04
4.45 ± 0.41
14.87 ± 2.00
0.49 ± 0.02
3.86 ± 0.48
11.77 ± 1.27*
0.52 ± 0.02
1.27 ± 0.20
1.24 ± 0.14
0.51 ± 0.03
1.00 ± 0.16
1.15 ± 0.02
0.47 ± 0.02
1.06 ± 0.11
1.14 ± 0.17
0.51 ± 0.02
1.24 ± 0.14
1.53 ± 0.20
The values are mean ± SEM of six individual fish.
a
HSI = liver weight × 100/body weight.
b
GSI = gonad weight × 100/body weight.
*
P < 0.05 indicates significant difference between exposure groups and the corresponding control.
significance was set as P < 0.05. All the data were expressed as the
mean ± standard error (SEM).
3. Results
3.1. Indices and production of eggs
No lethality was observed in zebrafish exposed to any
concentration of 2,4-DCP, nor were there effects on hepatic:somatic index (HSI) in either males or females (Table 2). The
gonad:somatic index (GSI) was 24% greater in males, but 35% less
in females exposed to 0.3 mg 2,4-DCP/L, than controls, respectively
(Table 2).
Cumulative production of eggs was significantly less in zebrafish
exposed to 0.3 mg 2,4-DCP/L (Table 3) and the effect was proportional to concentration of 2,4-DCP. Total content of protein and
diameter of eggs were not different in zebrafish exposed to any
of the concentrations of 2,4-DCP when compared to the controls
(Table 3).
Fig. 1. Concentrations of testosterone (T) and 17␤-estradiol (E2) in the culture
medium of H295R cells after exposure to 2,4-dichlorophenol (2,4-DCP) (0, 0.1, 0.3
and 1.0 mg/L) for 48 h. The values represent mean ± SEM of three replicate samples.
*P < 0.05 and **P < 0.01 indicate significant difference between exposure groups and
the corresponding control.
3.2. Concentrations of hormones
3.3. mRNA expression of genes
Exposure to 2,4-DCP caused statistically significant effects on
hormone production by H295R cells (Fig. 1). Concentrations of T
were not significantly changed in medium of H295R cells exposed
to 2,4-DCP (Fig. 1). Concentrations of E2 were significantly less than
those in the controls by 20%, 33% or 57% in the medium of cells
exposed to 0.1, 0.3 or 1.0 mg 2,4-DCP/L, respectively (Fig. 1).
Exposure to 2,4-DCP affected concentrations of hormones in
male and female zebrafish (Fig. 2). In males, concentrations of T
were significantly greater than that of the controls by 98%, 88%
or 136% when exposed to 0.03, 0.1 or 0.3 mg 2,4-DCP/L, respectively (Fig. 2A). In females, 2,4-DCP did not affect concentrations
of T in plasma (Fig. 2A). Concentrations of E2 in plasma of male
zebrafish were 13%, 38%, and 111% greater than those of controls, in increasing order of the three concentrations tested, but
the effect was statistically significant only in zebrafish exposed to
0.3 mg 2,4-DCP/L (Fig. 2B). In females, concentrations of E2 were
35%, 53%, and 71% less than those of control fish when exposed to
the three concentrations of 2,4-DCP, but the difference was statistically significant only for females exposed to 0.1 or 0.3 mg 2,4-DCP/L
(Fig. 2B).
Exposure to 2,4-DCP affected expression of mRNA for genes
involved in steroidogenesis (Fig. 3). After 48 h exposure, mRNA
expression of StAR and 17ˇHSD genes were significantly greater
by 1.8- and 1.6-fold in H295R cells exposed to 1.0 mg 2,4-DCP/L,
respectively (Fig. 3). Expression of CYP11A mRNA was significantly
up-regulated by 2.8- and 3.0-fold and expression of 3ˇHSD mRNA
was up-regulated by 1.6- and 2.0-fold in H295R cells exposed to
0.3 or 1.0 mg 2,4-DCP/L (Fig. 3). Expression of CYP17 mRNA was
significantly up-regulated by 2.2- and 2.3-fold when exposed to
0.3 or 1.0 mg 2,4-DCP/L (Fig. 3). Expression of CYP19 mRNA was
down-regulated by 2,4-DCP in a concentration-dependent manner
(Fig. 3).
Profiles of mRNA expression of genes involved in the HPG axis
and liver were affected by exposure to 2,4-DCP (Fig. 4). In the brain,
exposure to 2,4-DCP caused up-regulation of FSH expression of
1.9- and 2.6-fold in females exposed to 0.1 and 0.3 mg 2,4-DCP/L
(Fig. 4A). In males, mRNA expression of FSH was 1.7- and 2.4-fold
greater in zebrafish exposed to 0.1 or 0.3 mg 2,4-DCP/L, respectively
(Fig. 4B). Expression of CYP19B mRNA was down-regulated in the
Table 3
Reproductive endpoints in female zebrafish after exposure to 2,4-DCP (0, 0.03, 0.1, 0.3 mg/L) for 21 days.
Gamete parameters
a
Egg production (eggs/mating pair/day)
Egg protein (␮g/egg)b
Egg diameter (mm)b
a
b
*
0
0.03
0.1
0.3
45.40 ± 2.25
0.35 ± 0.01
1.15 ± 0.01
38.24 ± 2.98
0.35 ± 0.00
1.16 ± 0.00
34.83 ± 4.15
0.34 ± 0.01
1.16 ± 0.00
30.29 ± 3.58*
0.33 ± 0.01
1.17 ± 0.01
Values represent the mean ± SEM of three replicate tanks.
Values represent the mean ± SEM of three replicate tanks (15 eggs per tank).
P < 0.05 indicates significant difference between exposure groups and the corresponding control.
Y. Ma et al. / Aquatic Toxicology 106–107 (2012) 173–181
177
and ptgs2 by 2.1- and 2.1-fold, respectively (Fig. 4C). Significant
up-regulation of gene transcription was observed for ER˛ and FSHR
in females exposed to 0.3 mg 2,4-DCP/L by 2.2- and 2.1-fold, respectively (Fig. 4C). ERˇ mRNA was significantly down-regulated in
females 2.1-fold exposed to 0.3 mg 2,4-DCP/L (Fig. 4C). In the testes,
expression of 3ˇHSD mRNA was significantly up-regulated 2.2fold in zebrafish exposed to 0.3 mg 2,4-DCP/L (Fig. 4D). Expression
of both CYP17 and 17ˇHSD mRNA was up-regulated 1.5-, 1.8-fold
and 2.6-, 2.6-fold in zebrafish exposed to 0.1 or 0.3 mg 2,4-DCP/L,
respectively (Fig. 4D). Transcriptions of mRNAs for CYP19A, ER˛,
FSHR and ptgs2 were significantly up-regulated 3.8-, 2.1-, 2.9- and
6.4-fold in males exposed to the greatest concentration of 2,4-DCP,
relative to that of the control (Fig. 4D). A small up-regulation of
gene transcription was observed for ERˇ (Fig. 4D), but the effect
was not statistically significant.
In the liver, mRNA expression of VTG1, VTG3, ER˛ and ERˇ genes
were examined (Fig. 4). In females, the response of mRNA expression of VTG1 (1.9-, 3.3-fold), VTG3 (1.8-, 2.8-fold), ER˛ (2.4-, 3.8-fold)
and ERˇ (2.0-, 2.5-fold) genes showed up-regulation in zebrafish
exposed to 0.1 or 0.3 mg 2,4-DCP/L (Fig. 4E). In males transcription of VTG1 and VTG3 in the liver of zebrafish exposed to 0.1 or
0.3 mg 2,4-DCP/L was up-regulated by 2.0-, 2.7-fold and 2.2-, 2.3fold, respectively (Fig. 4F). Exposure to 2,4-DCP also increased the
mRNA expression of both ER˛ and ERˇ (Fig. 4F).
3.4. Hatching rates
Fig. 2. Plasma concentrations of T (A) and E2 (B) in zebrafish upon exposure to 2,4dichlorophenol (2,4-DCP) (0, 0.03, 0.1 or 0.3 mg/L) for 21 days. The values represent
mean ± SEM of eight individual fish from the three replicate tanks (2 fish pooled
as one replicate sample). *P < 0.05 and **P < 0.01 indicate significant differences
between treatments and the corresponding control.
females and up-regulated in the males when they were exposed to
2,4-DCP (Fig. 4A and B).
Expression of 3ˇHSD, CYP17, 17ˇHSD, CYP19A, ER˛, ERˇ, FSHR
and ptgs2 mRNAs were examined in gonads (Fig. 4). In the ovary,
exposure of females to 0.1 or 0.3 mg 2,4-DCP/L caused downregulation of CYP19A mRNA by 2.3- and 2.6-fold, respectively
Fig. 3. Gene expression profiles after exposure to 2,4-dichlorophenol (2,4-DCP) for
48 h in H295R cells. Values represent the mean ± SEM of three replicate samples.
Gene expressions were expressed as fold change relative to control. *P < 0.05 and
**P < 0.01 indicate significant difference between exposure groups and the corresponding control.
In the F1 generation, over 80% of the control embryos hatched
successfully 4 dpf (Fig. 5) and there was no significantly malformations observed in any of the treatments (data not shown). The rate
of hatching was reduced 14% and 22% when zebrafish were exposed
to 0.1 or 0.3 mg 2,4-DCP/L, respectively compare with the control
(Fig. 5). The embryos that did not hatch within 4 dpf were dead,
which suggests that maternal exposure to 2,4-DCP was the cause
of the effects, and perhaps due to toxicity of transferred toxicants
in the eggs and decreased gamete quality.
4. Discussion
H295R cells have been used for rapid screening of potential
disruption of steroidogenic pathways as well as determining mechanisms of action of toxicants (Sanderson et al., 2000; Hilscherova
et al., 2004; Zhang et al., 2005). However, few studies have compared the results of the in vitro H295R steroidogenesis assay with
the results of in vivo investigations (Han et al., 2010; Ji et al., 2010).
In the present study, H295R cells were used to assess effects on
transcription of specific genes involved in steroidogenesis. Those
results were then used to understand the effects of 2,4-DCP on
reproduction of zebrafish.
Measurement of steroid hormones released into the medium
has been suggested to be the most integrative, functional endpoint
and has been established previously in H295R cells (Hecker et al.,
2006). In the present study, lesser production of E2 was observed
after exposure of H295R cells to 2,4-DCP, and this was accompanied by down-regulation of mRNA expression of the aromatase
(CYP19) gene. CYP19 catalyzes the final and rate-limiting step in
conversion of androgen to estrogen and thus the down-regulation
of CYP19 gene expression could explain the lesser production of E2
in the presence of 2,4-DCP. This is consistent with those of previous
reports that down-regulation of expression of CYP19 mRNA could
result in less synthesis of E2 by H295R cells. For example, model
chemicals known to alter steroid metabolism, such as a binary mixture of forskolin and aminoglutethimide down-regulated mRNA
expression of CYP19 gene in H295R cells, which resulted in less production of E2 (Gracia et al., 2006). Exposure of H295R cells to PCP or
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Y. Ma et al. / Aquatic Toxicology 106–107 (2012) 173–181
Fig. 4. Gene expression profiles in zebrafish after exposure to 2,4-DCP (0, 0.03, 0.1 and 0.3 mg/L) for 21 days. Female brain (A); male brain (B); female gonad (C); male gonad
(D); female liver (E); male liver (F). Values represent the mean ± SEM of four individual fish from three replicate tanks. Gene expressions were expressed as fold change
relative to control. *P < 0.05 and **P < 0.01 indicate significant difference between exposure groups and the corresponding control.
2,4,6-TCP also resulted in less synthesis of E2 and down-regulation
of mRNA expression of CYP19 (Ma et al., 2011).
Synthesis of T in H295R cells was not significantly altered by
2,4-DCP, although mRNA expressions of genes coding for enzymes
important in steroidogenesis, including StAR, CYP11A, 3ˇHSD,
CYP17, and 17ˇHSD, were up-regulated. The gene product of CYP17
is important in the formation of androgens. One explanation of the
observed results is that up-regulation of StAR, 3ˇHSD, CYP11A and
CYP17 mRNA expressions might stimulate basal biosynthesis of
cortisol and aldosterone (Liu et al., 2010b). Therefore, measurement
of other steroid hormones would be helpful to understand the
overall effects on steroidogenesis. Furthermore, production of
steroids is a complex process with multiple sensitive control
points, so the genes within the steroidogenesis pathway are not
transcribed to the same extent and simple statistical correlations
between mRNA expression and hormone production should not
be expected (Gracia et al., 2007).
Both concentrations of hormones in plasma of zebrafish, and
transcription of genes in the HPG axis and liver were further investigated. Exposure to 2,4-DCP significantly altered concentrations of
Y. Ma et al. / Aquatic Toxicology 106–107 (2012) 173–181
Fig. 5. The hatching rate in offspring after maternal exposure to 2,4-dichlorophenol
(2,4-DCP) (0, 0.03, 0.1 or 0.3 mg/L) for 21 days. The values represent the mean ± SEM
of three replicate tanks (100 embryos per tank). *P < 0.05 and **P < 0.01 indicate
significant differences between treatments and the corresponding control.
hormones in plasma. Steroidogenic genes such as 3ˇHSD, 17ˇHSD
and CYP17 encode enzymes that participate in the synthesis of
testosterone. CYP19 is the terminal enzyme in the steroidogenic
pathway, which converts testosterone into estradiol, and the mRNA
level of CYP19 is well correlated with activity of aromatase (Trant
et al., 2001). In female zebrafish exposed to 2,4-DCP, production of
T was not affected, but lesser concentrations of E2 were observed
in plasma. The fact that concentrations of T in plasma were not
affected might be partially due to the unchanged mRNA expression
of 3ˇHSD, 17ˇHSD and CYP17, which was observed in zebrafish.
However, down-regulation of CYP19A mRNA was observed in the
ovary, which was consistent with the lesser concentrations of E2 in
plasma. In females, mRNA expression of CYP19A was significantly
down-regulated in the ovary while mRNA expression of CYP19B was
not significantly changed in brain. This result is consistent with the
results of a previous study that showed that CYP19A and CYP19B
present distinct profiles of mRNA expression in ovaries and brain
in zebrafish (Chiang et al., 2001).Concentrations of both T and E2
were greater in plasma of male zebrafish exposed to 2,4-DCP. The
greater production of T could be partially due to up-regulation of
mRNA expression of 3ˇHSD, 17ˇHSD, and CYP17 genes. Greater
concentrations of E2 in plasma of males were accompanied with
up-regulation of mRNA expression of CYP19A and CYP19B in the
testes and brain, respectively. These results can be explained by
increased conversion of testosterone to estradiol. In vertebrates
steroidogenic enzymes are critical for the production of androgens;
up-regulation of the mRNA expression of genes coding for these
enzymes can increase the efficiency of testosterone and estradiol
synthesis. In a previous study, 2,4,6-tribromophenol (TBP) exposure to zebrafish significantly increased T and E2 in males and
up-regulation of the mRNA expression of CYP17 and CYP19A genes
in testes (Deng et al., 2010).
Alternatively, mRNA expressions of the genes for FSH and FSHR
were up-regulated in both male and female zebrafish exposed to
2,4-DCP. FSH induces production of sex steroids and gametogenesis by first binding to their specific receptors (FSHR) in the gonads
of vertebrates (Kumar et al., 2001). Thus, up-regulation of transcription of these genes would be expected to result in greater
gamatogenesis and secretion of sex hormones by the testes. Exposure to 2,4-DCP resulted in a greater GSI in males and greater
synthesis of the steroid hormones, T and E2, changes which were
consistent with up-regulation of FSH mRNA expression in brain.
This observation would suggest promotion of gamatogenesis in
males. Treatment with either T or E2 resulted in significantly lesser
concentrations of FSH in plasma of coho salmon (Oncorhynchus
179
kisutch) (Dickey and Swanson, 1998). Alternatively, mRNA expression of FSHR was up-regulated in the ovary following inhibition
of aromatase and E2 production in fathead minnows by fadrozole
(Villeneuve et al., 2009). These results agree with the hypothesis
of a negative E2 feedback in the pituitary and hypothalamic. In
female zebrafish, the greater expression of FSH and FSHR mRNA
fish exposed to 2,4-DCP might be negative feedback regulation
in response to the reduction of plasma E2. It is possible that
up-regulation of expression of FSH and FSHR mRNA indicates compensation to less production of E2 production in female exposed to
2,4-DCP.
In teleosts, synthesis of prostaglandins (PGs) by follicles is
catalyzed by the enzyme cyclooxygenase (COX) and mediates maturation of oocytes and ovulation (Sorbera et al., 2001; Lister and
Van Der Kraak, 2008). The genome of the zebrafish contains one
COX-1 gene (ptgs1) and two functional COX-2 genes (ptgs2a and
ptgs2b) (Fujimoria et al., 2011). mRNA expression of the Ptgs2
gene can be induced when male fathead minnows are exposed to
ethinyl estradiol (EE2) (Garcia-Reyero et al., 2009). When female
zebrafish were exposed to environmentally relevant doses of di(2-ethylhexyl)-phthalate (DEHP), mRNA expression of the ptgs2
gene was significantly decreased following inhibition of ovulation
and egg production (Carnevali et al., 2010). In this study, down
regulation of expression of ptgs2 mRNA in ovary accompanied by
significantly less production of eggs by zebrafish exposed to 2,4DCP. However, the role of PG in reproduction of male fish remains
unclear.
Synthesis of vitellogenin (VTG) in the liver is usually responsive
to stimulation by estrogenic chemicals which bind to specific ERs
to activate the transcription of VTG gene. Hence, the greater mRNA
expression of hepatic VTG1 and VTG3 observed in males exposed to
2,4-DCP, could either be result of increased E2 in plasma, or could
suggest that 2,4-DCP was acting as a xenoestrogen. It has previously reported that exposure to TBP resulted in significantly greater
concentrations of E2 in plasma and a subsequent up-regulation of
mRNA expression of VTG in the liver of male zebrafish (Deng et al.,
2010). However, transcription of VTG was up-regulated although
concentrations of E2 in plasma of females were less. In a yeast
two-hybrid assay, 2,4-DCP has been shown to be estrogenic by initiating the binding of ER and ER responsive element (ERE) (Nishihara
et al., 2000), and promote proliferation of cells in human breast
tumor cells (MCF-7) (Jones et al., 1998). A plausible explanation
of this observation is that 2,4-DCP might be a direct-acting ERagonist, to stimulate vitellogenin synthesis. This is consistent with
the observation that both ER˛ and ERˇ were transcriptionally significantly up-regulated in liver of zebrafish exposed to 2,4-DCP. A
previous study showed that4-nonylphenol (NP) stimulated synthesis of VTG in rainbow trout, whereas concentrations of E2 in plasma
was less and the authors speculated that NP directly act on ERs and
stimulated VTG synthesis in the liver (Harris et al., 2001). In an
in vitro study, using cultured hepatocytes of male rainbow trout,
it was shown that NP induced mRNA expression of VTG and ER˛
genes (Jobling and Sumpter, 1993). Exposure of female rare minnow (G. rarus) to 2,4-DCP caused alternation of ERs in liver (Zhang
et al., 2008). Therefore, up-regulation of expression of ER mRNAs in
the liver appears to be a common response and is responsible for
chemical-induced VTG synthesis in the liver.
In vitro assays could provide screening and classification of
potential endocrine disrupting of chemicals, however, it is still a
challenge to validate results from in vitro assays to predict these
responses in vivo and to relate the changes at the level of gene
expression to developmental and reproductive performance of the
animal (Segner et al., 2003). In the study reported on here, effects of
2,4-DCP on mRNA expression of the steroidogenic genes CYP17 and
17ˇHSD was similar between H295R and the male fish gonad, but
not in females. A previous study using the anti-androgen flutamide,
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Y. Ma et al. / Aquatic Toxicology 106–107 (2012) 173–181
showed up-regulation of genes coding for enzymes involved in
androgen biosynthesis, such as CYP17, which implied an inhibitory
action on androgen negative feedback pathways (Filby et al., 2007).
Although mRNA expression of genes in H295R cells and male
fish gonad was similar, concentrations of T were not significantly
changed and production of E2 was less in H295R cells, meanwhile,
both concentrations of T and E2 were higher in males exposed
to 2,4-DCP. The results of the study reported here indicate that
responses in the in vitro H295R cell-based steroidogenesis assay
cannot predict the effects of a chemical on synthesis of steroid
hormones in fish gonad. The mechanisms that resulted in greater
production of T and E2 are not well-known. A possible explanation is that 2,4-DCP might have anti-androgenic activity in vivo
as in vitro (Li et al., 2010). In male fish, exposed to 2,4-DCP had
insufficient androgenic signaling, as a result, mRNA expression of
steroidogenic genes was stimulated through feedback, resulting
in greater production of endogenous T and thus greater production of E2 due to increased availability of T as well as greater
mRNA expression of CYP19A, and consequent increases in liver
VTG expression. However, the observation that mRNA expression
of steroidogenic genes in ovary did not change, but concentrations of both T and E2 in plasma were less cannot be as readily
explained by an anti-androgenic mode of action. The mechanisms
underlying the gender dependent responses in zebrafish are not
clear.
Few studies have been conducted to compare endocrine disruption of toxicants on steroidogenesis between H295R cells and fish. A
comparative study of effects of fadrozole and fenarimol onsteroidogenesis in H295R cells and ovary explants from fathead minnow
found that responses differed between the two assays (Villeneuve
et al., 2007b). The authors suggested that the differences might be
due to different mechanisms underlying the effects and also some
of these differences might be attributed to differences between
steroidogenic pathways in H295R cells and those in gonads of fish.
It should also be noted that fish and mammalian steroidogenic
enzymes might also differ in specificity for some chemicals (Baker,
2001). Consequently, some steroidogenic processes are likely to be
regulated differently between fish and mammals, which could also
result in different sensitivities and/or responses to endocrine disruptors (Villeneuve et al., 2007b). In a whole organism, endocrine
disruptors interfere with multiple endocrine functions, which are
regulated by feedback mechanisms. In vitro assays do not take
account of absorption, distribution, metabolism, or excretion. Thus,
results obtained from in vitro assays could not be used to predict the
risk of endocrine disruptors in an intact organism. Hence, in vitro
tests using H295R cells are suitable in a Tier I screening assay for
environmental risk assessment, and further in vivo studies will be
needed to investigate the impact of endocrine disruptors on the
steroidogenic pathway at the whole-organism level.
In summary, the complexity of endocrine function in animals
makes it difficult to rely on in vitro data alone to predict the
endocrine disrupting properties of xenobiotics. Therefore, it is useful to compare effects of chemicals in vitro and in vivo. The results
of this study were meant to be comparative, rather than predictive
and the results should not be used to predict effects under environmental conditions. 2,4-DCP exposure affects the production of sex
hormones by changing the expression of several key steroidogenic
genes in H295R cell line in vitro. In the vivo experiment, waterborne
exposure of 2,4-DCP to zebrafish could modulate the expressions of
genes in the HPG-axis and disrupt steroidogenesis in a systematic
manner, which in turn, can cause adverse effects on reproductive
success in the offspring. It has also been suggested that profiles
of expression of genes in the HPG axis provides a useful tool to
both illuminate chemical-induced modes of action and to quantitatively evaluate chemical-induced adverse effects (Ankley et al.,
2009).
Acknowledgements
This work was supported by Chinese Academy of Sciences
(KZCX2-YW-Q02-05), the National Nature Science Foundation of
China (no. 20890113), and the State Key Laboratory of Freshwater
Ecology and Biotechnology (2008FBZ10). The research was supported, in part, by a Discovery Grant from the National Science
and Engineering Research Council of Canada (Project # 32641507) and a grant from the Western Economic Diversification Canada
(Project # 6578 and 6807). Prof. Giesy was supported by the Canada
Research Chair program, an at large Chair Professorship at the
Department of Biology and Chemistry and State Key Laboratory
in Marine Pollution, City University of Hong Kong, The Einstein
Professor Program of the Chinese Academy of Sciences and the Visiting Professor Program of King Saud University. The authors also
wish to thank the two anonymous reviewers for their constructive
comments.
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