Disruption of endocrine function in in vitro H295R cell-based and in in vivo assay in zebrafish by 2,4-dichlorophenol

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: [email protected] (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

174

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

176

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

178

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,

180

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. References Ankley, G.T., Jensen, K.M., Durhan, E.J., Makynen, E.A., Butterworth, B.C., Kahl, M.D., Villeneuve, D.L., Linnum, A., Gray, L.E., Cardon, M., Wilson, V.S., 2005. Effects of two fungicides with multiple modes of action on reproductive endocrine function in the fathead minnow (Pimephales promelas). Toxicol. Sci. 86, 300– 308. Ankley, G.T., Bencic, D.C., Cavallin, J.E., Jensen, K.M., Kahl, M.D., Makynen, E.A., Martinovic, D., Mueller, N.D., Wehmas, L.C., Villeneuve, D.L., 2009. Dynamic nature of alterations in the endocrine system of fathead minnows exposed to the fungicide prochloraz. Toxicol. Sci. 112, 344–353. Baker, M.E., 2001. Hydroxysteroid dehydrogenases: ancient and modern regulators of adrenal and sex steroid action. Mol. Cell. Endocrinol. 175, 1–4. Brillas, E., Calpe, J.C., Casado, J., 2000. Mineralization of 2,4-D by advanced electrochemical oxidation processes. Water Res. 34, 2253–2262. Carnevali, O., Tosti, L., Speciale, C., Peng, C., Zhu, Y., Maradonna, F., 2010. DEHP impairs zebrafish reproduction by affecting critical factors in oogenesis. PLoS ONE 5, e10201. Chiang, E.F., Yan, Y.L., Yann, G.G., Postlethwait, J., Chung, B.C., 2001. Two Cyp19 (P450 aromatase) genes on duplicated zebrafish chromosomes are expressed in ovary or brain. Mol. Biol. Evol. 18, 542–550. Chiron, S., Minero, C., Evione, D., 2007. Occurrence of 2,4-dichlorophenol and of 2,4dichloro-6-nitrophenol in the Rhône River Delta (Southern France). Environ. Sci. Technol. 41, 3127–3133. Deng, J., Liu, C., Yu, L., Zhou, B., 2010. Chronic exposure to environmental levels of tribromophenol impairs zebrafish reproduction. Toxicol. Appl. Pharmacol. 243, 87–95. Dickey, J.T., Swanson, P., 1998. Effects of sex steroids on gonadotropin (FSH and LH) regulation in coho salmon (Oncorhynchus kisutch). Mol. Endocrinol. 21, 291–306. Ding, L., Murphy, M.B., He, Y., Xu, Y., Yeung, L.W.Y., Wang, J., Zhou, B., Lam, P.K.S., Wu, R.S.S., Giesy, J.P., 2007. Effects of brominated flame retardants and brominated dioxins on steroidogenesis in H295R human adrenocortical carcinoma cell line. Environ. Toxicol. Chem. 26, 764–772. Filby, A.L., Tyler, C.R., 2007. Appropriate housekeeping genes for use in expression profiling the effects of environmental estrogens in fish. BMC Mol. Biol. 8, 10. Filby, A.L., Thorpe, K.L., Maack, G., Tyler, C.R., 2007. Gene expression profiles revealing the mechanisms of anti-androgen-and estrogen-induced feminization in fish. Aquat. Toxicol. 81, 219–231. Fujimoria, C., Ogiwaraa, K., Hagiwaraa, A., Rajapakseb, S., Kimuraa, A., Takahashi, T., 2011. Expression of cyclooxygenase-2 and prostaglandin receptor EP4b mRNA in the ovary of the medaka fish, Oryzias latipes: possible involvement in ovulation. Mol. Cell. Endocrinol. 332, 67–77. Gao, J., Liu, L., Liu, X., Zhou, H., Huang, S., Wang, Z., 2008. Levels and spatial distribution of chlorophenols 2,4-dichlorophenol, 2,4,6-trichlorophenol, and pentachlorophenol in surface water of China. Chemosphere 71, 1181–1187. Garcia-Reyero, N., Kroll, K.J., Liu, L., Orlando, E.F., Watanabe, K.H., Sepúlveda, M.S., Villeneuve, D.L., Perkins, E.J., Ankley, G.T., Denslow, N.D., 2009. Gene expression responses in male fathead minnows exposed to binary mixtures of an estrogen and antiestrogen. BMC Genomics 10, 308. Gracia, T., Hilscherova, K., Jones, P.D., Newsted, J.L., Zhang, X., Hecker, M., Higley, E.B., Sanderson, J.T., Yu, R.M.K., Wu, R.S.S., Giesy, J.P., 2006. The H295R system for evaluation of endocrine-disrupting effects. Ecotoxicol. Environ. Saf. 65, 293– 305. Gracia, T., Hilscherova, K., Jones, P.D., Newsted, J.L., Higley, E.B., Zhang, X., Hecker, M., Murphy, M.B., Yu, R.M.K., Lam, P.K.S., Wu, R.S.S., Giesy, J.P., 2007. Modulation of steroidogenic gene expression and hormone production of H295R cells by pharmaceuticals and other environmentally active compounds. Toxicol. Appl. Pharmacol. 225, 142–153. Han, S.Y., Choi, K., Kim, J.K., Ji, K.H., Kim, S.M., Ahn, B.W., Yun, J.H., Choi, K.H., Khim, J.S., Zhang, X., Giesy, J.P., 2010. Endocrine disruption and consequences of chronic exposure to ibuprofen in Japanese medaka (Oryzias latipes) and freshwater cladocerans Daphnia magna and Moina macrocopa. Aquat. Toxicol. 98, 256–264.

Y. Ma et al. / Aquatic Toxicology 106–107 (2012) 173–181 Harris, C.A., Santos, E.M., Janbakhsh, A., Pottinger, T.G., Tyler, C.R., Sumpter, J.P., 2001. Nonylphenol affects gonadotropin levels in the pituitary gland and plasma of female rainbow trout. Environ. Sci. Technol. 35, 2909–2916. Harvey, P.W., Everett, D.J., Springall, C.J., 2007. Adrenal toxicology: a strategy for assessment of functional toxicity to the adrenal cortex and steroidogenesis. J. Appl. Toxicol. 27, 103–115. He, Y., Murphy, M.B., Yu, R.M., Lam, M.H., Hecker, M., Giesy, J.P., Wu, R.S., Lam, P.K., 2008. Effects of 20 PBDE metabolites on steroidogenesis in the H295R cell line. Toxicol. Lett. 176, 230–238. Hecker, M., Newsted, J.L., Murphy, M.B., Higley, E.B., Jones, P.D., Wu, R., Giesy, J.P., 2006. Human adrenocarcinoma (H295R) cells for rapid in vitro determination of effects on steroidgenesis: hormone production. Toxicol. Appl. Pharmacol. 217, 114–124. Hilscherova, K., Jones, P.D., Gracia, T., Newsted, J.L., Zhang, X., Sanderson, J.T., Yu, R.M., Wu, R.S., Giesy, J.P., 2004. Assessment of the effects of chemicals on the expression of ten steroidogenic genes in the H295R cell line using real-time PCR. Toxicol. Sci. 81, 78–89. House, W.A., Leach, D., Long, J.L., Cranwell, P., Smith, C., Bharwaj, L., Meharg, A., Ryland, G., Orr, D.O., Wright, J., 1997. Micro-organic compounds in the Humber rivers. Sci. Total Environ. 194–195, 357–371. Ji, K., Choi, K., Lee, S., Park, S., Khim, J.S., Jo, E.H., Choic, K., Zhang, X., Giesy, J.P., 2010. Effects of sulfathiazole, oxytetracycline and chlortetracycline on steroidogenesis in the human adrenocarcinoma (H295R) cell line and freshwater fish Oryzias latipes. J. Hazard. Mater. 182, 494–502. Jobling, S., Sumpter, J.P., 1993. Detergent components in sewage effluent are weakly oestrogenic to fish: an in vitro study using rainbow trout (Oncorhynchus mykiss) hepatocytes. Aquat. Toxicol. 27, 361–372. Jones, P.A., Baker, V.A., Irwin, A.J.E., Earl, L.K., 1998. Interpretation of the in vitro proliferation response of MCF-7 cells to potential oestrogens and non- oestrogenic substances. Toxicol. In Vitro 12, 373–382. Kumar, R., Ijiri, S., Trant, J., 2001. Molecular biology of channel catfish gonadotropin receptors. 2. Complementary DNA cloning, functional expression, and seasonal gene expression of the follicle-stimulating hormone receptor. Biol. Reprod. 65, 710–717. Li, L.A., Wang, P.W., 2005. PCB126 induces differential changes in androgen, cortisol, and aldosterone biosynthesis in human adrenocortical H295R cells. Toxicol. Sci. 85, 530–540. Li, J., Ma, M., Wang, Z., 2010. In vitro profiling of endocrine disrupting effects of phenols. Toxicol. In Vitro 24, 210–217. Lister, A.L., Van Der Kraak, G., 2008. An investigation into the role of prostaglandins in zebrafish oocyte maturation and ovulation. Gen. Comp. Endocrinol. 159, 46–57. Liu, C., Yu, L., Deng, J., Lam, P.K., Wu, R.S., Zhou, B., 2009. Waterborne exposure to fluorotelomer alcohol 6:2 FTOH alters plasma sex hormone and gene transcription in the hypothalamic-pituitary-gonadal (HPG) axis of zebrafish. Aquat. Toxicol. 93, 131–137. Liu, C., Deng, J., Yu, L., Ramesh, M., Zhou, B., 2010a. Endocrine disruption and reproductive impairment in zebrafish by exposure to 8:2 fluorotelomer alcohol. Aquat. Toxicol. 96, 70–76. Liu, C., Zhang, X., Chang, H., Jones, P., Wiseman, S., Naile, J., Hecker, M., Giesy, J.P., Zhou, B., 2010b. Effects of fluorotelomer alcohol 8:2 FTOH on steroidogenesis in H295R cells: targeting the cAMP signaling cascade. Toxicol. Appl. Pharmacol. 247, 222–228. Livak, K.J., Schmittgen, T.D., 2001. Analysis of relative gene expression data using real-time quantitative PCR and the 2−CT method. Methods 25, 402– 408. Ma, Y., Liu, C., Lam, P.K., Wu, R.S., Giesy, J.P., Hecker, M., Zhang, X., Zhou, B., 2011. Modulation of steroidogenic gene expression and hormone synthesis in H295R cells exposed to PCP and TCP. Toxicology 282, 146–153. Nagahama, Y., Yamashita, M., 2008. Regulation of oocyte maturation in fish. Dev. Growth Differ. 50, S195–S219.

181

Nishihara, T., Nishikawa, J., Kanayama, T., Dakeyama, F., Saito, K., Imagawa, M., Takatori, S., Kitagawa, Y., Hori, S., Utsumi, H., 2000. Estrogenic activities of 517 chemicals by yeast two-hybrid assay. J. Health Sci. 46, 282–298. Ohlsson, A., Ullerås, E., Oskarsson, A., 2009. A biphasic effect of the fungicide prochloraz on aldosterone, but not cortisol, secretion in human adrenal H295R cells-underlying mechanisms. Toxicol. Lett. 191, 174–180. Ramamoorthy, S., Ramamoorthy, S., 1997. Chlorinated Organic Compounds in the Environment. Regulatory and Monitoring Assessment. Lewis Publishers, Boca Ration. Richards, J.S., 1994. Hormonal control of gene-expression in the ovary. Endocr. Rev. 15, 725–751. Sanderson, J.T., Seinen, W., Giesy, J.P., van den Berg, M., 2000. 2-Chloro-s-triazine herbicides induce aromatase (CYP19) activity in H295R human adrenocortical carcinoma cells: a novel mechanism for estrogenicity? Toxicol. Sci. 54, 121–127. Segner, H., Navas, J.M., Schafers, C., Wenzel, A., 2003. Potencies of estrogenic compounds in in vitro screening assays and in life cycle tests with zebrafish in vivo. Ecotoxicol. Environ. Saf. 54, 315–322. Sofikitis, N., Giotitsas, N., Tsounapi, P., Baltogiannis, D., Giannakis, D., Pardalidis, N., 2008. Hormonal regulation of spermatogenesis and spermiogenesis. Steroid Biochem. Mol. Biol. 109, 323–330. Sorbera, L.A., Asturiano, J.F., Carrillo, M., Zanuy, S., 2001. Effects of polyunsaturated fatty acids and prostaglandins on oocyte maturation in a marine teleost, the European sea bass (Dicentrarchus labrax). Biol. Reprod. 64, 382–389. Stringer, R., Johnston, P., 2001. Chlorine and the Environment: An Overview of the Chlorine Industry. Kluwer Academic Publishers, Dordercht, Boston. Trant, J.M., Gavasso, S., Ackers, J., Chung, B.C., Place, A.R., 2001. Developmental expression of cytochrome P450 aromatase genes (CYP19a and CYP19b) in zebrafish fry (Danio rerio). J. Exp. Zool. 290, 475–483. Uren-Webster, T.M., Lewis, C., Filby, A.L., Paull, G.C., Santos, E.M., 2010. Mechanisms of toxicity of di(2-ethylhexyl) phthalate on the reproductive health of male zebrafish. Aquat. Toxicol. 99, 360–369. Villeneuve, D.L., Larkin, P., Knoebl, I., Miracle, A.L., Kahl, M.D., Jensen, K.M., Makynen, E.A., Durhan, E.J., Carter, B.J., Denslow, N.D., Ankley, G., 2007a. A graphical systems model to facilitate hypothesis-driven ecotoxicogenomics research on the teleost brain–pituitary–gonadal axis. Environ. Sci. Technol. 41, 321–330. Villeneuve, D.L., Ankley, G.T., Makynen, E.A., Blake, L.S., Greene, K.J., Higley, E.B., Newsted, J.L., Giesy, J.P., Hecker, M., 2007b. Comparison of fathead minnow ovary explant and H295R cell-based steroidogenesis assays for identifying endocrineactive chemicals. Ecotoxicol. Environ. Saf. 68, 20–32. Villeneuve, D.L., Mueller, N.D., Martinovic, D., Makynen, E.A., Kahl, M.D., Jensen, K.M., Durhan, E.J., Cavallin, J.E., Bencic, D., Ankley, G.T., 2009. Direct effects, compensation, and recovery in female fathead minnows exposed to a model aromatase inhibitor. Environ. Health Perspect. 117, 624–631. Zhang, X., Hecker, M., Jones, P.D., Newsted, J., Au, D., Kong, R., Wu, R.S.S., Giesy, J.P., 2008a. Responses of the medaka HPG axis PCR array and reproduction to prochloraz and ketoconazole. Environ. Sci. Technol. 42, 6762–6769. Zhang, X., Hecker, M., Park, J.W., Tompsett, A.R., Newsted, J., Nakayama, K., Jones, P.D., Au, D., Kong, R., Wu, R.S., Giesy, J.P., 2008b. Real-time PCR array to study effects of chemicals on the hypothalamic-pituitary-gonadal axis of the Japanese medaka. Aquat. Toxicol. 88, 173–182. Zhang, X., Yu, R.M., Jones, P.D., Lam, G.K., Newsted, J.L., Gracia, T., Hecker, M., Hilscherova, K., Sanderson, T., Wu, R.S., Giesy, J.P., 2005. Quantitative RT-PCR methods for evaluating toxicant-induced effects on steroidogenesis using the H295R cell line. Environ. Sci. Technol. 39, 2777–2785. Zhang, X., Zha, J., Li, W., Yang, L., Wang, Z., 2008. Effects of 2,4-dichlorophenol on the expression of vitellogenin and estrogen receptor genes and physiology impairments in Chinese rare minnow (Gobiocypris Rarus). Environ. Toxicol. 23, 694–701. Zona, R., Solar, S., Gehringer, P., 2002. Degradation of 2,4-dichlorophenoxyacetic acid by ionizing radiation: influence of oxygen concentration. Water Res. 36, 1369–1374.