Modulations in androgen and estrogen mediating genes and testicular response in male goldfish exposed to bisphenol A (original) (raw)

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South Bohemia Research Center of Aquaculture and Biodiversity of Hydrocenoses, Faculty of Fisheries and Protection of Waters, University of South Bohemia in České Budějovice, Vodňany, Czech Republic

South Bohemia Research Center of Aquaculture and Biodiversity of Hydrocenoses, Faculty of Fisheries and Protection of Waters, University of South Bohemia in České Budějovice, Vodňany, Czech Republic

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Department of Biological Sciences, University of Calgary, Calgary, Alberta, Canada

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Sayyed Mohammad Hadi Alavi

South Bohemia Research Center of Aquaculture and Biodiversity of Hydrocenoses, Faculty of Fisheries and Protection of Waters, University of South Bohemia in České Budějovice, Vodňany, Czech Republic

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Department of Biological Sciences, University of Calgary, Calgary, Alberta, Canada

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South Bohemia Research Center of Aquaculture and Biodiversity of Hydrocenoses, Faculty of Fisheries and Protection of Waters, University of South Bohemia in České Budějovice, Vodňany, Czech Republic

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26 April 2012

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Azadeh Hatef, Ava Zare, Sayyed Mohammad Hadi Alavi, Hamid R. Habibi, Otomar Linhart, Modulations in androgen and estrogen mediating genes and testicular response in male goldfish exposed to bisphenol A, Environmental Toxicology and Chemistry, Volume 31, Issue 9, 1 September 2012, Pages 2069–2077, https://doi.org/10.1002/etc.1919
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Abstract

Adverse effects of bisphenol A (BPA) on reproductive physiology were studied in male goldfish (Carassius auratus) exposed to nominal environmentally relevant concentrations (0.2 and 20 µg/L) for up to 90 d. Transcriptions of various reproductive genes were measured in brain, liver, and testis to investigate the BPA modes of action. Volume, density, total number, motility, and velocity of sperm were measured to assess testicular function. At 0.2 µg/L, BPA reduced steroidogenetic acute regulatory protein and increased estrogen receptors (ERs) messenger RNA (mRNA) transcript (ERβ1 in liver and ERβ2 in testis) after 90 d. At 20 µg/L, BPA increased mRNA transcript of androgen receptor in testis, brain‐ and testis‐specific aromatase, and vitellogenin in liver after 90, 30, 60, and 60 d, respectively. Transcripts of ERs mRNA were increased after 30 to 60 d at 20 µg/L BPA; increase in ERβ1 mRNA was observed in testis after 7 d. Total number, volume, and motility of sperm were decreased in males exposed to 0.2 and 20 µg/L BPA, whereas sperm density and velocity were only reduced at 20 µg/L BPA. The results support the hypothesis that BPA may exert both anti‐androgenic and estrogenic effects, depending on concentration, leading to diminished sperm quality. The findings provide a framework for better understanding of the mechanisms mediating adverse reproductive actions of BPA observed in different parts of the world. Environ. Toxicol. Chem. 2012; 31: 2069–2077. © 2012 SETAC

INTRODUCTION

Bisphenol A (BPA) is widely used for production of polymers such as polycarbonates, epoxy, phenolic resins, polyesters, and polyacrylates to manufacture polycarbonate plastic products, resins lining cans, dental sealants, and food industry products, such as containers, plastic wrappers, and inner coatings for cans 1, 2. Studies in mammals demonstrated that BPA exerts both estrogenic and anti‐androgenic effects, depending on concentration and exposure time 2–4. In fish, most studies showed estrogenic effects of BPA such as stimulation of vitellogenin (Vtg) and transcription alternations of specific genes in different organs 5, 6. However, the adverse effects of BPA at environmentally relevant concentrations are still unclear on reproductive performance in fish. Moreover, BPA molecular modes of action are still unknown because transcription alternations have been studied in a few numbers of genes and have not been compared between organs involved in reproduction, including brain, liver, and gonad.

Synthesis of Vtg in the liver is induced by 17β‐estradiol (E2) 7. Although the Vtg gene is present in the liver of both sexes, it is only elevated in females because of endogenous estrogens. In males, Vtg is low and only becomes detectable when they are exposed to estrogen or xenoestrogens. Therefore, Vtg has been used for the identification of estrogenic‐like contaminants 5, 8, 9. Induction of Vtg gene and its synthesis has been shown in a concentration‐ and time‐dependent manner in different fish species exposed to BPA at usually high concentrations (higher than 100 µg/L) after two to four weeks of exposure 10–13. The present study investigated whether lower BPA concentrations approaching environmental levels alter Vtg messenger RNA (mRNA) transcript in liver.

The physiological functions of estrogen receptors (ERs) in E2‐mediated induction of Vtg have been investigated in goldfish 14, 15, demonstrating direct involvement of ERβ subtypes in Vtg synthesis as well as the induction of ERα. Several studies show modulations in both ERα and ERβ in mammals administered by BPA 16–19, but BPA affinity to ERs is unknown in fish. Additionally, BPA effects on aromatase P450 in brain (CYP19b) and testis (CYP19a), which mediates conversion of androgens into estrogens 20, 21, were studied in the present study. Results of ERs and P450 aromatase provide valuable information about the mechanism in which normal E2‐mediated Vtg synthesis is disrupted by BPA.

Decrease of androgens has been previously reported in fish exposed to BPA at low (0.6–11 µg/L) 22 or high (1,000 µg/L) 13 concentrations, but mechanisms of action are still unknown. Studies show a decrease of androgens mediated by luteinizing hormone secretion from pituitary via androgen receptor (AR) antagonist 19, 23, 24. Sharpe et al. 25 suggested that a decrease in androgens occurs via down‐regulation of steroidogenic acute regulatory protein (StAR) that delivers cholesterol to the inner mitochondrial membrane in male goldfish exposed to β‐sitosterol. In the present study, AR and StAR modulations were measured for understanding of BPA effects in decreasing androgens.

Two studies show adverse effects of BPA on sperm production and motility in fish 22, 26. These studies revealed delays in spermiation and decrease in sperm quality in males exposed to BPA at environmentally relevant concentrations. In the present study, sperm quality was also investigated after 90 d exposure to BPA to investigate the correlation between mRNA transcript of reproductive genes and change in sperm quality. Goldfish were exposed to BPA at 0.2 and 20 µg/L, which overlaps with a range of concentrations reported in the European rivers 3.

MATERIAL AND METHODS

Experimental design

Mature male goldfish (Carassius auratus) (2‐ to 3‐year‐old) were obtained from Hluboka fish farm, Hluboka nad Vltavou, Czech Republic, at the beginning of the spawning season. The health status of fish was first checked by random selection of some individuals. Neither parasite nor disease was detected in their gills or mucus. Then, they were acclimatized and distributed into the aquaria. All groups, including control and BPA‐treated groups, were under similar condition; each group was kept in 70‐L aquaria and maintained under a 12‐h light/12‐h dark photoperiod. To be closer to physiological condition, water in the aquaria was not heated, and temperature was measured as 18 and 22°C at the beginning and end of the experiment, respectively. Water dissolved oxygen and pH were checked three times per week before feeding and were 6.0 mg/L (5.7–6.2) and 7.5 (7.3–7.8), respectively. The fish were fed once per day with commercial food (ZZN Vodnany).

Fish were exposed to BPA (Sigma‐Aldrich) in May 2009. Duplicate tanks of adult male goldfish were exposed to nominal 0.2 and 20 µg/L BPA and solvent control for a period of 90 d. Dimethyl sulfoxide (Sigma‐Aldrich) was used to dissolve BPA, and its final concentration was 0.001% in each aquarium. A control group without dimethylsulfoxide was also studied, but the results did not show any significant difference from solvent control (see Tables 1 and 2, and Fig. 1). Therefore, only the data from solvent control corresponding to each sampling time were used as reference for comparison with respective treatments. Every other day, 80% of the exposure solution was renewed.

Table 1.

Body mass (g), total length (cm), gonadosomatic index (GSI, %), and hepatosomatic index (HSI, %) of male goldfish (Carassius auratus L.) exposed to different concentrations of bisphenol A (BPA) sampled at 7, 15, 30, 60, and 90 d after exposurea,b

a

Solvent control group was exposed to 0.001% dimethylsulfoxide (DMSO).

b

No significant differences were observed between groups at each sampling time (p > 0.05).

Table 1.

Body mass (g), total length (cm), gonadosomatic index (GSI, %), and hepatosomatic index (HSI, %) of male goldfish (Carassius auratus L.) exposed to different concentrations of bisphenol A (BPA) sampled at 7, 15, 30, 60, and 90 d after exposurea,b

a

Solvent control group was exposed to 0.001% dimethylsulfoxide (DMSO).

b

No significant differences were observed between groups at each sampling time (p > 0.05).

Table 2.

Sperm volume, sperm density, and total number of sperm cells in male goldfish (Carassius auratus L.) exposed to different concentrations of bisphenol A (BPA) after 90 d exposurea

a

Solvent control group was exposed to 0.001% dimethylsulfoxide (DMSO) (n = 8).

b

At each column, values with different superscripts are significantly different (p < 0.05).

Table 2.

Sperm volume, sperm density, and total number of sperm cells in male goldfish (Carassius auratus L.) exposed to different concentrations of bisphenol A (BPA) after 90 d exposurea

a

Solvent control group was exposed to 0.001% dimethylsulfoxide (DMSO) (n = 8).

b

At each column, values with different superscripts are significantly different (p < 0.05).

Figure 1.

Sperm motility (A) and velocity (B) in male goldfish exposed to bisphenol A (BPA). The fish were exposed to nominal 0.2 and 20 µg/L BPA and sperm samples were collected at 90 d following exposure. Data represents mean ± SEM. Values with different superscripts are significantly different for sperm volume, density and total number of spermatozoa (n = 8, p < 0.05). In case of sperm motility and velocity, values with different superscripts are significantly different at each time post activation (n = 8, p < 0.05).

Thirty individuals were introduced to each aquarium to collect five samples at 7, 15, 30, and 60 d and 10 samples at 90 d after exposure. Sample size differences in the present study were attributable to the existence of female individuals. This was because of no sexual character to recognize males and females. Samples were collected from at least three individuals at each sampling time. At each sampling time, fish were first anesthetized in 2‐phenoxyethanol dissolved in water (0.3 ml/L). At 90 d after exposure, sperm collection for quality assessment (methods outlined later) was performed before sampling of organs. Each male was individually weighted (± 0.1 g) and measured for total length (± 1 mm). Fish were then sacrificed by cutting spinal cord; testis and liver were removed and weighed (± 0.01 g) for determination of gonadosomatic index (GSI) (= 100 × gonad wt/total fish wt) and hepatosomatic index (HSI) (= 100 × liver wt/total fish wt). Samples were immediately frozen in liquid nitrogen and kept in an RNAse‐free tube at −80°C until use.

RNA extraction and complementary DNA synthesis

Total RNA was extracted from each tissue sample of male fish, using Trizol Reagent (Invitrogen, Cat. No. 15596‐018), following the manufacturer's instructions. Total RNA concentration was estimated from absorbance at 260 nm (A260 nm, Nanodrop, USA), and RNA quality was verified by A260 nm/A280 nm ratios between 1.8 and 2 and A230 nm/A260 nm ratios greater than 2. Complementary DNA was synthesized from 4 µg total RNA of each sample using Moloney Murine Leukemia Virus Reverse Transcriptase (M‐MLV) (Invitrogen, Cat No. 28025‐013) and oligo (dT)18 primer (Promega) following the manufacturer's instructions. Briefly, 2 µl Oligo (dT)18 primer (500 µg/ml) was added to each sample and the reaction mixture was heated to 70°C for 10 min and then quickly chilled to 4°C. After cooling, 4 µl of 5× first‐strand buffer, 2 µl 100 mM dithiothreitol, 0.4 µl deoxyribonucleotide (dNTP; 100 mM; cat num:dNTP‐01,UBI Life Science) and 0.7 µl M‐MLV (200 U/µl) were added to a total volume of 18 µl. The reaction mixture was then incubated at 25°C for 10 min and at 37°C for 50 min using an iQ cycler. By heating at 70°C for 15 min, the reaction was deactivated. Each 18‐µl reaction was diluted threefold in nuclease‐free water and used as a template for quantitative real‐time polymerase chain reaction (PCR) assay.

Quantitative real‐time PCR

The iCycler iQ Real‐time PCR Detection System (Bio‐Rad Laboratories) was used for evaluating gene expression level with the following condition per reaction: 1 µl diluted complementary DNA, 0.26 µM each primer, 12.5 µl SYBR Green PCR Master Mix(Qiagen), and ultrapure distilled water (Invitrogen) to a total volume of 25 µl. Specific primers were used to detect expression of different genes (Table 3) 15. The thermal profile for all reactions was 3 min at 95°C and then 45 cycles of 10 s at 95°C, 20 s appropriate annealing temperature and 72°C. The specificity of the amplified product in the quantitative PCR assay was determined by analyzing the melting curve to discriminate target amplicon from primer dimer or other nonspecific products. A single melt curve was observed for each primer set in all quantitative PCR reactions. Each individual sample was run in triplicate, and the mean threshold cycles (as determined by the linear portion of the fluorescence absorbance curve) were used for the final calculation. The mRNA expression levels were normalized to the expression level of glyceraldehyde‐3‐ phosphate dehydrogenase mRNA, using the standard 2−ΔΔCt method.

Table 3.

Primer sequences and amplification (annealing) programs used for measurements of selected genes using qRT‐PCR

qRT‐PCR = quantitative real‐time polymerase chain reaction; StAR = steroidogenic acute regulatory protein; CYP19a = gonad aromatase; CYP19b = brain aromatase; AR = androgen receptor; ERα, ERβ1, and ERβ2 = estrogen receptor subtypes; Vtg = vitellogenin; GAPDH = glyceraldehyde 3‐phosphate dehydrogenase.

Table 3.

Primer sequences and amplification (annealing) programs used for measurements of selected genes using qRT‐PCR

qRT‐PCR = quantitative real‐time polymerase chain reaction; StAR = steroidogenic acute regulatory protein; CYP19a = gonad aromatase; CYP19b = brain aromatase; AR = androgen receptor; ERα, ERβ1, and ERβ2 = estrogen receptor subtypes; Vtg = vitellogenin; GAPDH = glyceraldehyde 3‐phosphate dehydrogenase.

Sperm quality assessment

Although our previous study showed the effects of BPA on sperm quality 22, confirming the decrease in sperm quality in the present study was important. Therefore, sperm quality was evaluated only once, following 90 d exposure. Sperm was collected from each individual by a gentle abdominal massage from the anterior portion of the testis toward the genital papilla and collected with plastic syringes and kept on ice (2−4°C) until analysis. Attention was focused on avoiding sperm contamination by urine, mucus, blood, and water.

Sperm volume and density were measured and expressed as milliliters and milliard of sperm per milliliter, respectively. Sperm density was measured by counting, following the method of Alavi et al. 27. Semen was diluted two times, each 100‐fold with 0.7% NaCl. Ten microliters diluted semen was placed onto a haemocytometer covered with a coverslip and left for 10 min to allow sperm sedimentation before 16 cells (0.1‐mm depth and 0.2 length) were counted. Total number of spermatozoa is recounted as spermatozoa concentration × sperm volume.

To evaluate the sperm motility traits (percentage of sperm motility and sperm velocity), sperm of each individual was directly activated in NaCl 50 mM, KCl 5 mM, Tris 20 mM, pH 8.5, 110 mOsmol/kg at ratio 1:1,000–2,000 (sperm:activation solution). To avoid sperm stickness into the slides, bovine serum albumin was added into the activation medium right before adding sperm at a final concentration of 0.1% w/v. Sperm motility was recorded using a CCD video camera (SONY DXC‐970MD) mounted on a darkfield microscope (Olympus BX50) supplied with a stroboscopic lamp and a DVD‐recorder (SONY DVO‐1000 MD). Then, the five successive frames were captured and analyzed with a micro image analyzer (Olympus Micro Image 4.0.1. for Windows) to measure sperm motility (%) and velocity (µm/s) based on successive position of sperm heads (for details see Hatef et al. 28). For each individual, three separate records were performed, and the mean of the values was used in statistical analysis. Data of sperm velocity presented in results are from only motile spermatozoa. For each record, sperm velocities were measured in 30 to 40 spermatozoa, and mean values were used in statistical analysis.

Statistical analysis

All data were analyzed using SPSS 9.0 for Windows, and each male was considered a replicate. Data for AR, ER subtypes, P450 aromatase, and Vtg were log transformed to meet assumptions of normality and homoscedasticity examined by Levene's test. Repeated‐measure analysis of variance (ANOVA; alpha at 0.05) was first used to understand the effects of main factors; BPA concentration, exposure time, and their interaction 24, 32. Results were summarized as _F_df, ddf P, where F is F value, df is degree of freedom, ddf is error of df, and p is p value. To understand the effects of BPA concentrations, the model was then revised into one‐way ANOVA models followed by Tukey's post‐hoc test at each sampling time (7, 15, 30, 60, or 90 d); either significant (ERβ1 and ERβ2 in liver) or nonsignificant (remained parameters) effects of main factors, and their interactions were observed. In case of sperm parameters, only one main factor was present (BPA concentration); therefore, one‐way analysis of variance was performed to understand the effects of BPA on sperm parameters. All data are presented as mean ± standard error of mean.

RESULTS

No mortality was observed in any treatments or sampling time. Body mass and total length remained unchanged between the BPA‐treated groups and controls (Table 1). Neither GSI nor HSI was affected by BPA concentration and exposure time (Tables 1 and 4).

Table 4.

Summary of statistics (Fdf, ddf) obtained from repeated‐measures analysis of variance (ANOVA) models used to study the effects of main factors (bisphenol A [BPA] concentration, exposure time, and their interaction) on reproductive performances in male goldfish (Carassius auratus L.)

*

p < 0.05.

**

p < 0.01.

***

p < 0.001.

GSI = gonadosomatic index; HIS = hepatosomatic index; StAR = steroidogenic acute regulatory protein; AR = androgen receptor; CYP19 = P450 aromatase; ERα, ERβ1, and ERβ2 = estrogen receptor subtypes; Vtg = vitellogenin.

Table 4.

Summary of statistics (Fdf, ddf) obtained from repeated‐measures analysis of variance (ANOVA) models used to study the effects of main factors (bisphenol A [BPA] concentration, exposure time, and their interaction) on reproductive performances in male goldfish (Carassius auratus L.)

*

p < 0.05.

**

p < 0.01.

***

p < 0.001.

GSI = gonadosomatic index; HIS = hepatosomatic index; StAR = steroidogenic acute regulatory protein; AR = androgen receptor; CYP19 = P450 aromatase; ERα, ERβ1, and ERβ2 = estrogen receptor subtypes; Vtg = vitellogenin.

StAR mRNA transcript

Exposure to BPA significantly influenced StAR mRNA transcript in the testis (Table 4). A decrease in StAR mRNA transcript was observed after 90 d exposure of males to 0.2 µg/L BPA (_F_2 = 6.3, p < 0.01, Fig. 2A).

Figure 2.

Mean ± SEM messenger RNA (mRNA) transcript of steroidogenic acute regulatory protein (StAR) (A) and androgen receptor (AR) (B) in testis of male goldfish exposed to bisphenol A (BPA). The fish were exposed to nominal 0.2 and 20 µg/L BPA and transcript abundance is expressed relative to that of solvent control (0.001% dimethylsulfoxide [DMSO]). At each exposure time, values with different superscripts are significantly different (n = 3–8, p < 0.05).

AR mRNA transcript

Transcript of AR mRNA in the testis was influenced by BPA concentration and exposure time (Table 4). The AR mRNA transcript was significantly lower in fish exposed to 0.2 µg/L BPA compared with those of 20 µg/L after 7 (_F_2 = 2.1) and 15 d (_F_2 = 2.5) exposure (p = 0.05), but the changes were not statistically different from those of solvent control (Fig. 2B). After 90 d exposure, AR mRNA transcript was increased at 20 µg/L BPA (_F_2 = 6.8, p < 0.01, Fig. 2B).

Aromatase mRNA transcript

Both BPA concentration and exposure time significantly influenced CYP19b and CYP19a mRNA transcript in the brain and testis (Table 4). Increase in CYP19b mRNA transcript was observed in fish exposed to 20 µg/L BPA after 30 (_F_2 = 6.0, p < 0.05) and 60 d (_F_2 = 11.4, p < 0.01) exposure (Fig. 3A). In the testis, CYP19a mRNA transcript was increased after 60 (_F_2 = 2.53, p < 0.05) and 90 d (_F_2 = 6.1, p < 0.01) in fish exposed to 20 µg/L BPA (Fig. 3B).

Figure 3.

Mean ± SEM messenger RNA (mRNA) transcript of P450 aromatase in brain (CYP19b) (A) and testis (CYP19a) (B) of male goldfish exposed to bisphenol A (BPA). The fish were exposed to nominal 0.2 and 20 µg/L BPA and transcript abundance is expressed relative to that of solvent control (0.001% dimethylsulfoxide [DMSO]). At each exposure time, values with different superscripts are significantly different (n = 3–8, p < 0.05).

ER subtypes mRNA transcript

Transcript of ERα mRNA was affected by BPA concentration in the brain and by exposure time in the liver (Table 4). Increase in the brain ERα mRNA transcript was observed after 30 (_F_2 = 8.1) and 60 d (_F_2 = 2.0) exposure to 20 µg/L BPA (p < 0.01, Fig. 4). In the testis and liver, the ERα mRNA transcript was unchanged (data are not shown).

Figure 4.

Mean ± SEM messenger RNA (mRNA) transcript of estrogen receptor (ERα) in the brain of male goldfish exposed to bisphenol A (BPA). The fish were exposed to nominal 0.2 and 20 µg/L BPA and transcript abundance is expressed relative to that of solvent control (0.001% dimethylsulfoxide [DMSO]). At each exposure time, values with different superscripts are significantly different (n = 3–8, p < 0.05).

Transcript of ERβ1 mRNA was affected by BPA concentration in the brain, liver, and testis, as well as exposure time in the brain and liver (Table 4). In the brain, increase in ERβ1 mRNA transcript was observed after 60 d exposure to 20 µg/L (_F_2 = 1.8, p < 0.05, Fig. 5A). In the testis, ERβ1 mRNA transcript was increased at all time points after exposure to 20 µg/L; 7 d (_F_2 = 3.9), 15 d (_F_2 = 4.9), 30 d (_F_2 = 3.6), 60 d (_F_2 = 1.8), and 90 d (_F_2 = 4.2) (p < 0.05, Fig. 5B). In the liver, ERβ1 mRNA transcript was increased at 60 d (_F_2 = 3.7, p < 0.05) and 90 d (_F_2 = 8.4, p < 0.01) in fish exposed to 20 µg/L and after 90 d exposure to 0.2 µg/L (_F_2 = 2.8, p < 0.05) (Fig. 5C).

Figure 5.

Mean ± SEM messenger RNA (mRNA) transcript of estrogen receptor (ERβ1) in brain (A), testis (B) and liver (C) of male goldfish exposed to bisphenol A (BPA). The fish were exposed to nominal 0.2 and 20 µg/L BPA and transcript abundance is expressed relative to that of solvent control (0.001% dimethylsulfoxide [DMSO]). At each exposure time, values with different superscripts are significantly different (n = 3–8, p < 0.05).

Both BPA concentration and exposure time affected ERβ2 mRNA transcript in the brain, liver, and testis (Table 4). Significant effect of BPA concentration and exposure time interaction was only observed in the liver (Table 4). In the brain, ERβ2 mRNA transcript was increased in fish exposed to 20 µg/L BPA after 30 d exposure (_F_2 = 7.0, p < 0.01) (Fig. 6A). In the testis, ERβ2 mRNA transcript was increased in fish exposed to 20 µg/L BPA after 60 d (_F_2 = 2.1, p < 0.05) and to 0.2 and 20 µg/L BPA after 90 d exposure (_F_2 = 15.7, p < 0.001, Fig. 6B). In liver, an increase in ERβ2 mRNA transcript was observed after 60 (_F_2 = 5.053) and 90 d (_F_2 = 7.7) exposure to 20 µg/L BPA (p < 0.01, Fig. 6C).

Figure 6.

Mean ± SEM messenger RNA (mRNA) transcript of estrogen receptor (ERβ2) in brain (A), testis (B) and liver (C) of male goldfish exposed to bisphenol A (BPA). The fish were exposed to nominal 0.2 and 20 µg/L BPA and transcript abundance is expressed relative to that of solvent control (0.001% dimethylsulfoxide [DMSO]). At each exposure time, values with different superscripts are significantly different (n = 3–8, p < 0.05).

Vtg mRNA transcript

The BPA concentration and exposure time influenced Vtg mRNA transcript in the liver (Table 4). An increase in Vtg mRNA transcript was observed after 60 (_F_2 = 4.3) and 90 d (_F_2 = 5.1) in fish exposed to 20 µg/L BPA (p < 0.05, Fig. 7).

Figure 7.

Mean ± SEM messenger RNA (mRNA) transcript of vitellogenin receptor (Vtg) in the liver of male goldfish exposed to bisphenol A (BPA). The fish were exposed to nominal 0.2 and 20 µg/L BPA and transcript abundance is expressed relative to that of solvent control (0.001% dimethylsulfoxide [DMSO]). At each exposure time, values with different superscripts are significantly different (n = 3–8, p < 0.05).

Sperm quality

Sperm volume (_F_2 = 4.7) and total number of spermatozoa (_F_2 = 3.2) were decreased in both groups exposed to 0.2 and 20 µg/L BPA (p < 0.05, Table 2). Sperm density was decreased in males exposed to 20 µg/L BPA (_F_2 = 2.3, p < 0.05, Table 2). Sperm motility was decreased in fish exposed to 20 µg/L (_F_2 = 20.6, p < 0.001) evaluated at 15 s post activation and in both 0.2 and 20 µg/L BPA evaluated at 30 s (_F_2 = 32.2, p < 0.001), 45 s (_F_2 = 11.2, p < 0.001), and 60 s (_F_2 = 5.2, p < 0.01) after activation (Fig. 1A). Sperm velocity showed a significant decrease in fish exposed to 20 µg/L BPA evaluated at 15 s (_F_2 = 10.5, p < 0.001), 30 s (_F_2 = 4.1, p < 0.05), 45 s (_F_2 = 3.3, p < 0.05), and 60 s (_F_2 = 6.9, p < 0.01) post activation (Fig. 1B).

DISCUSSION

For better understanding of the BPA mechanism of toxicity leading to reproductive endocrine disruption, the present study was performed to investigate correlation between sperm quality and mRNA transcript levels of genes involved in testicular function in goldfish. This study was carried out as a follow‐up to our previous observation that exposure to BPA results in a decline in sperm quality associated with a decrease in androgens 22. The results demonstrate changes in mRNA transcripts of sex steroids, Vtg, and steroidogenic‐mediating genes encoding steroidogenic enzymes in goldfish after exposure to 0.2 and 20 µg/L BPA. At 0.2 µg/L, StAR mRNA transcript was reduced in the testis and mRNA transcript of ERβ1 in the liver, and ERβ2 in the testis were increased after 90 d exposure (Table 5). At 20 µg/L, transcript of P450 aromatase, ERs, Vtg, and AR mRNA were increased after 30, 60, or 90 d exposure (Table 5). The BPA concentrations examined in the present study are relevant to the levels observed in the European rivers 3, 6.

Table 5.

Summary of the bisphenol A (BPA) effects on messenger RNA (mRNA) transcripts of estrogen and androgen receptors, vitellogenein, and steroidogenesis mediating genes in male goldfish (Carassius auratus L.)

NS = no significant effect was observed; StAR = steroidogenic acute regulatory protein; AR = androgen receptor; CYP19a = testis aromatase; CYP19b = brain aromatase; ERα, ERβ1 and ERβ2 = estrogen receptor subtypes; Vtg = vitellogenin.

Table 5.

Summary of the bisphenol A (BPA) effects on messenger RNA (mRNA) transcripts of estrogen and androgen receptors, vitellogenein, and steroidogenesis mediating genes in male goldfish (Carassius auratus L.)

NS = no significant effect was observed; StAR = steroidogenic acute regulatory protein; AR = androgen receptor; CYP19a = testis aromatase; CYP19b = brain aromatase; ERα, ERβ1 and ERβ2 = estrogen receptor subtypes; Vtg = vitellogenin.

No mortality was observed in BPA‐treated groups, indicating that the examined BPA concentrations were sublethal. The GSI and HSI were unchanged, which are consistent with previous works on goldfish, common carp, and medaka exposed to 0.6 to 11, 1 to 1,000 and 3,120 µg/L BPA for 30, 15, and 21 d, respectively 11, 13, 22, 29. Similarly, testis weight was also found to remain unchanged in mammals 18, 19, 30. These findings suggest that GSI and HSI are not good indicators for screening adverse effects of BPA on fish reproduction.

The increase in AR mRNA transcript observed at 20 µg/L was concurrent with our previous results showing reduction in androgen production in goldfish exposed to 11 µg/L BPA 22. A decrease in StAR mRNA transcript was observed at each time point examined, but the significant decrease was only observed after 90 d exposure to 0.2 µg/L BPA. This is because of high variations resulting from inter‐animal variability as well as sample size. The observed decrease of StAR, in part, might be responsible for the reported decrease in androgen synthesis due to disruption of cholesterol transport to mitochondria. Similarly, Sharpe et al. 25 reported a decrease in StAR and androgen concentration in goldfish exposed to β‐sitosterol after five months of exposure. Moreover, inhibition of androgen production interferes with a decrease of luteinizing hormone secretion in mammals via inhibition of 17α‐hydroxylase 19, 24. These observations collectively suggest dose‐dependent BPA modes of action to disrupt androgen biosynthesis in testis. At low doses, inhibition of androgen synthesis occurs via disruption of pituitary and testis functions to inhibit luteinizing hormone secretion or through a decrease in substrate availability mediated by StAR. At high doses, BPA acts through competitive binding to AR, which in turn suppresses luteinizing hormone function to stimulate androgen biosynthesis 19, 24.

Exposure to 0.2 µg/L BPA stimulated mRNA transcripts of ERβ1 in liver and ERβ2 in testis after 90 d, but these modulations did not lead to an increase in Vtg mRNA transcript. The observed increase might be attributable to the affinity of BPA to ERs. Exposure to 20 µg/L BPA stimulated all three subtypes of ER in brain, liver, or testis, which led to an increase in Vtg mRNA transcript. The observed increase in ERs by BPA is consistent with the presumed estrogenic mimicking action 14, 15. Similarly, BPA increased ERα and ERβ1 mRNA transcript in tilapia 31. Our results are also consistent with the report that in the brain, liver, and testis of a hermaphroditic fish; exposure to 600 µg/L BPA for 4 d increased ERα expression, whereas ERβ mRNA transcript remained unchanged 32. Differences in BPA‐induced ERs mRNA transcript might be attributable to species variability, maturity stage of fish, and concentration or period of exposure 4. In the rat brain, BPA stimulated ERβ mRNA transcript without affecting that of ERα 19. In the rat testis, BPA decreased and increased ERβ and ERα mRNA transcripts, respectively 19. One study in male rats demonstrated an increase in both ERβ and ERα mRNA transcript in the brain, probably because of high concentration of BPA (100 µg/kg body wt per day). Contributing factors that may explain observed variability in ERs' response to BPA include differences in exposure time, concentration, and mode of administration as well as differences in affinity of BPA to various ERs 15, 17, 19, 33. The present results demonstrate tissue‐specific transcription response to BPA in addition to seasonally related changes in estrogenic activity reported previously 19, 34, 35.

Observed increases in CYP19a and CYP19b mRNA transcripts in fish exposed to 20 µg/L BPA are in agreement with previous works on medaka and zebrafish exposed to a high concentration of BPA 36–38. These suggest that P450 aromatase family are targets of BPA and can be used as a biomarker for screening reproductive disruption.

The observed stimulation of Vtg mRNA transcript indicates estrogenic property of BPA at 20 µg/L in the present study, which encompasses environmental level. Stimulation of Vtg synthesis in liver has been frequently reported in response to BPA, for instance, in common carp 13, goldfish 22, swordtail 39, fathead minnow 10, 40, medaka 11, 12, and rainbow trout 41. The reported effective concentrations for inducing Vtg production are different among these studies that might be addressed to BPA uptake rates, species‐specific ER binding affinities, fish species, age, maturity stage, and exposure period. However, all of these observations suggest estrogenic activity of BPA at high concentration, which is similar to other estrogens such as (E2 and 17β‐ethynylestradiol) or estrogen‐mimicking compounds such as nonylphenol and phthalate 40, 42.

Observed decrease of sperm quality is in agreement with Lahnsteiner et al. 26 and Hatef et al. 22, when brown trout and goldfish were exposed to 1.75 to 2.40 and 0.6 to 11 µg/L BPA, respectively. Bisphenol A may modulate sperm maturation via alternations in sex steroid biosynthesis in testis 22. In mammals, a concentration‐dependent reduction in sperm motility has been reported after the administration of BPA to male rats 24, 43. Meeker et al. 44 reported a relationship between sperm motility and urinary BPA in men examined in an infertility clinic. However, the mechanisms of adverse effects of BPA on sperm quality and sperm motility are still unknown. Further studies are required to investigate modes of action of BPA on sperm maturation by studying sperm maturation–inducing hormone, intracellular cyclic adenosine monophosphate and pH in sperm, initial adenosine triphosphate content, plasma membrane potential for sperm activation 45–47, as well as sperm cytoskeletal proteins, DNA damage, and oxidative stress 30, 43, 44.

A model is proposed for the adverse effects of BPA on reproductive function, based on the results of the present study and previous findings (Table 5, Fig. 8). At low concentrations, BPAs decrease androgens via modulation in testicular steroidogenesis functions, particularly because of a decrease in substrate availability mediated by suppression StAR. These changes could lead to a decrease in sperm quality. At high concentrations, BPA induces Vtg production in the liver mediated by stimulation in mRNA transcript of ERs and P450 aromatase, which converts androgens to estrogen. Under physiological conditions, all three ERs contribute in Vtg 15. However, ERα expression is induced by E2 through activation of ERβ subtypes, and ERβ1 is responsible for maintaining the ERα level in the liver. The present study suggests that BPA induces Vtg induction, particularly mediated by ERβ subtypes, because ERα expression remained unchanged in the liver.

Figure 8.

A model summary for the effects of bisphenol A (BPA) on reproductive mediated genes in goldfish at low and high environmentally relevant concentration. ER = estrogen receptor; CYP19b = brain aromatase; StAR = steroidogenic acute regulatory protein.

In conclusion, the present study shows that the BPA‐mediated actions likely involve a number of mechanisms, including changes in ER expression, aromatase activity, and potential pathways interfering with androgen‐sensitive response. The BPA mode of action is dose‐dependent, but it interferes with testicular functions and reduces the sperm quality. Further studies are needed to understand the potential roles of sex steroids and their receptors in sperm maturation or signaling cascades required for sperm activation.

Acknowledgements

The authors declare no conflict of interests. The present study was supported by Grant Agency of the Czech Republic 523/09/1793 and P503/12/1834 to SMHA, CENAKVA CZ.1.05/2.1.00/01.0024 and Grant Agency of the University of South Bohemia 047/2010/Z and 046/2010/Z to OL, and Natural Sciences and Engineering Research Council of Canada (NSERC) grants to HRH. We are grateful to M. Pečená for her technical assistance during sampling. All animals were handled according to §17 odst. 1 zakona No. 246/119 Sb, Ministry of Agriculture of the Czech Republic.

REFERENCES

1

Oehlmann

J

,

Oetken

M

,

Schulte‐Oehlmann

U

.

2008

.

A critical evaluation of the environmental risk assessment for plasticizers in the freshwater environment in Europe, with special emphasis on bisphenol A and endocrine disruption

.

Environ Res

198

:

140

–

149

.

2

Vandenberg

LN

,

Maffini

MV

,

Sonnenschein

C

,

Rubin

BS

,

Soto

AM

.

2009

.

Bisphenol‐A and the great divide: A review of controversies in the field of endocrine disruption

.

Endocrine Rev

30

:

75

–

95

.

3

Crain

DA

,

Eriksen

M

,

Iguchi

T

,

Jobling

S

,

Laufer

H

,

LeBlanc

GA

,

Guillette

LJ

Jr.

2007

.

An ecological assessment of bisphenol‐A: Evidence from comparative biology

.

Reprod Toxicol

24

:

225

–

239

.

4

Wetherill

YB

,

Akingbemi

BT

,

Kanno

J

,

McLachlan

JA

,

Nadal

A

,

Sonnenschein

C

,

Watson

CS

,

Zoeller

RT

,

Belcher

SM

.

2007

.

In vitro molecular mechanisms of bisphenol A action: Review

.

Reprod Toxicol

24

:

178

–

198

.

5

Jeffries

KM

,

Jackson

LJ

,

Ikonomou

MG

,

Habibi

HR

.

2010

.

Presence of natural and anthropogenic organic contaminants and potential fish health impacts along two river gradients in Alberta, Canada

.

Environ Toxicol Chem

29

:

2379

–

2388

.

6

Kang

JH

,

Aasi

D

,

Katayama

Y

.

2007

.

Bishenol A in the aquatic environment and its endocrine disruptive effects on aquatic organisms

.

Crit Rev Toxicol

37

:

607

–

625

.

7

Finn

RN

.

2007

.

Vertebrate yolk complexes and the functional implications of phosvitins and other subdomains in vitellogenins

.

Biol Reprod

76

:

926

–

935

.

8

Sumpter

JP

,

Jobling

S

.

1995

.

Vitellogenesis as a biomarker for estrogenic contamination of the aquatic environment

.

Environ Health Perspect

103

:

173

–

178

.

9

Jeffries

KM

,

Nelson

ER

,

Jackson

LJ

,

Habibi

HR

.

2008

.

Basin‐wide impacts of compounds with estrogen‐like activity on longnose dace (Rhinichthys cataractae) in two Prairie Rivers of Alberta, Canada

.

Environ Toxicol Chem

27

:

2042

–

2052

.

10

Sohoni

P

,

Tyler

CR

,

Hurd

K

,

Caunter

J

,

Hetheridge

M

,

Williams

T

.

2001

.

Reproductive effects of long‐term exposure to bisphenol A in the fathead minnow (Pimephales promelas)

.

Environ Sci Technol

35

:

2917

–

2925

.

11

Kang

IJ

,

Yokota

H

,

Oshima

Y

,

Tsuruda

Y

,

Oe

T

,

Imada

N

,

Tadokoro

H

,

Honjo

T

.

2002

.

Effects of bisphenol A on the reproduction of Japanese medaka (Oryzias latipes)

.

Environ Toxicol Chem

21

:

2394

–

2400

.

12

Tabata

A

,

Watanabe

N

,

Yamamoto

I

,

Ohnishi

Y

,

Itoh

M

,

Kamei

T

,

Magara

Y

,

Terao

Y

.

2004

.

The effect of bisphenol A and chlorinated derivatives of bisphenol A on the level of serum vitellogenin in Japanese medaka (Oryzias latipes)

.

Water Sci Technol

50

:

125

–

132

.

13

Mandich

A

,

Bottero

S

,

Benfenati

E

,

Cevasco

A

,

Erratico

C

,

Maggioni

S

,

Massari

A

,

Pedemonte

F

,

Viganň

L

.

2007

.

In vivo exposure of carp to graded concentrations of bisphenol A

.

Gen Comp Endocrinol

153

:

15

–

24

.

14

Nelson

ER

,

Wiehler

WB

,

Cole

WC

,

Habibi

HR

.

2007

.

Homologous regulation of estrogen receptor subtypes in goldfish

.

Mol Reprod Dev

74

:

1105

–

1112

.

15

Nelson

ER

,

Habibi

HR

.

2010

.

Functional significance of nuclear estrogen receptor subtypes in the liver of goldfish

.

Endocrinology

151

:

1668

–

1676

.

16

Gould

JC

,

Leonard

LS

,

Maness

SC

,

Wagner

BL

,

Conner

K

,

Zacharewski

T

,

Safe

S

,

McDonnell

DP

,

Gaido

KW

.

1998

.

Bisphenol A interacts with the estrogen receptor‐alpha in a distinct manner from estradiol

.

Mol Cell Endocrinol

142

:

203

–

214

.

17

Kuiper

GG

,

Lemmen

JG

,

Carlsson

B

,

Corton

JC

,

Safe

SH

,

Van Der Saag

PT

,

van der Burg

B

,

Gustafsson

JA

.

1998

.

Interaction of estrogenic chemicals and phytoestrogens with estrogen receptor‐beta

.

Endocrinology

139

:

4252

–

4263

.

18

Takao

T

,

Nanamiya

W

,

Pournajafi Nazarloo

H

,

Matsumoto

R

,

Asaba

K

,

Hashimoto

K

.

2003

.

Exposure to the environmental estrogen bisphenol A differentially modulated estrogen receptor‐α and –β immunoreactivity and mRNA in male mouse testis

.

Life Sci

72

:

1159

–

1169

.

19

Akingbemi

BT

,

Sottas

CM

,

Koulova

AI

,

Klinefelter

GR

,

Hardy

MP

.

2004

.

Inhibition of testicular steroidogenesis by the xenoestrogen bisphenol A is associated with reduced pituitary luteinizing hormone secretion and decreased steroidogenic enzyme gene expression in rat Leydig cells

.

Endocrinology

145

:

592

–

603

.

20

Tchoudakova

A

,

Callard

GV

.

1998

.

Identification of multiple CYP19 genes encoding different cytochrome P450 aromatase isozymes in brain and ovary

.

Endocrinology

139

:

2179

–

2189

.

21

Barney

ML

,

Patil

JG

,

Gunasekera

RM

,

Carter

CHJ

.

2008

.

Distinct cytochrome P450 aromatase isoforms in the common carp (Cyprinus carpio): Sexual dimorphism and onset of ontogenic expression

.

Gen Comp Endocrinol

156

:

499

–

508

.

22

Hatef

A

,

Alavi

SMH

,

Abdulfatah

A

,

Fontaine

P

,

Rodina

M

,

Linhart

O

.

2012

.

Adverse effects of Bisphenol A on reproductive physiology in male goldfish at environmentally relevant concentrations

.

Ecotoxicol Environ Saf

76

:

56

–

62

.

23

Xu

LC

,

Sun

H

,

Chen

JF

,

Bian

Q

,

Qian

J

,

Song

L

,

Wang

XR

.

2005

.

Evaluation of androgen receptor transcriptional activities of bisphenol A, octylphenol and nonylphenol in vitro

.

Toxicology

216

:

197

–

203

.

24

Salian

S

,

Doshi

T

,

Vanage

G

.

2009

.

Neonatal exposure of male rats to Bisphenol A impairs fertility and expression of sertoli cell junctional proteins in the testis

.

Toxicology

265

:

56

–

67

.

25

Sharpe

RL

,

Woodhouse

A

,

Moon

TW

,

Trudeau

VL

,

MacLatchy

DL

.

2007

.

β‐Sitosterol and 17β‐estradiol alter gonadal steroidogenic acute regulatory protein (StAR) expression in goldfish, Carassius auratus

.

Gen Comp Endocrinol

151

:

34

–

41

.

26

Lahnsteiner

F

,

Berger

B

,

Kletzl

M

,

Weismann

T

.

2005

.

Effect of bisphenol A on maturation and quality of semen and eggs in the brown trout, Salmo trutta f. fario

.

Aquat Toxicol

75

:

213

–

224

.

27

Alavi

SMH

,

Rodina

M

,

Policar

T

,

Linhart

O

.

2009

.

Relationships between semen characteristics and body size in Barbus barbus L. (Teleostei: Cyprinidae) and effects of ions and osmolality on sperm motility

.

Comp Biochem Physiol

153A

:

430

–

437

.

28

Hatef

A

,

Alavi

SMH

,

Linhartova

Z

,

Rodina

M

,

Policar

T

,

Linhart

O

.

2010

.

In vitro effects of Bisphenol A on sperm motility characteristics in Perca fluviatilis L. (Percidae; Teleostei)

.

J Appl Ichthyol

26

:

696

–

701

.

29

Ishibashi

H

,

Watanabe

N

,

Matsumura

N

,

Hirano

M

,

Nagao

Y

,

Shiratsuchi

H

,

Kohra

S

,

Yoshihara

S

,

Arizono

K

.

2005

.

Toxicity to early life stages and an estrogenic effect of a bisphenol A metabolite, 4‐methyl‐2, 4‐bis (4‐hydroxyphenyl) pent‐1‐ene on the medaka (Oryzias latipes)

.

Life Sci

77

:

2643

–

2655

.

30

Peknicova

J

,

Kyselova

V

,

Buckiova

D

,

Boubelik

M

.

2002

.

Effect of an endocrine Disruptor on mammalian fertility: Application of monoclonal antibodies against sperm proteins as markers for testing sperm damage

.

Am J Reprod Immunol

47

:

311

–

318

.

31

Huang

W

,

Zhang

Y

,

Jia

X

,

Ma

X

,

Li

S

,

Liu

Y

,

Zhu

P

,

Lu

D

,

Zhao

H

,

Luo

W

,

Yi

S

,

Liu

X

,

Lin

H

.

2010

.

Distinct expression of three estrogen receptors in response to bisphenol A and nonylphenol in male Nile tilapias (Oreochromis niloticus)

.

Fish Physiol Biochem

36

:

237

–

249

.

32

Seo

JS

,

Lee

YM

,

Jung

SO

,

Kim

IC

,

Yoon

YD

,

Lee

JS

.

2006

.

Nonylphenol modulates expression of androgen receptor and estrogen receptor genes differently in gender types of the hermaphroditic fish Rivulus marmoratus

.

Biochem Biophys Res Commun

346

:

213

–

223

.

33

Matthews

JB

,

Twomey

K

,

Zacharewski

TR

.

2001

.

In vitro and in vivo interactions of bisphenol A and its metabolite, bisphenol A glucuronide, with estrogen receptors α and β

.

Chem Res Toxicol

14

:

149

–

157

.

34

Choi

CY

,

Habibi

HR

.

2003

.

Molecular cloning of estrogen receptor α and expression pattern of estrogen receptor subtypes in male and female goldfish

.

Mol Cell Endocrinol

204

:

169

–

177

.

35

Nelson

ER

,

Habibi

HR

.

2008

.

Seasonal‐related homologous regulation of goldfish liver Estrogen receptor subtypes

.

Cybium

32

:

248

–

249

.

36

Lee

YM

,

Seo

JS

,

Kim

IC

,

Yoon

YD

,

Lee

JS

.

2006

.

Endocrine disrupting chemicals (bisphenol A, 4‐nonylphenol, 4‐tert‐octylphenol) modulate expression of two distinct cytochrome P450 aromatase genes differently in gender types of the hermaphroditic fish, Rivulus marmoratus

.

Biochem Biophys Res Commun

345

:

894

–

903

.

37

Min

J

,

Lee

SK

,

Gu

MB

.

2003

.

Effects of endocrine disrupting chemicals on distinct expression patterns of estrogen receptor, cytochrome P450 aromatase and p53 genes in Oryzias latipes liver

.

J Biochem Mol Toxicol

17

:

272

–

277

.

38

Kishida

M

,

McLellan

M

,

Miranda

JA

,

Callard

GV

.

2001

.

Estrogen and xenoestrogens upregulate the brain aromatase isoform (P450aromB) and perturb markers of early development in zebrafish (Danio rerio)

.

Comp Biochem Physiol

129B

:

261

–

268

.

39

Kwak

HJ

,

Bae

MO

,

Lee

MH

,

Lee

YS

,

Lee

BJ

,

Kang

KS

,

Chae

CL

,

Sung

HJ

,

Shin

JS

,

Kim

JH

,

Mar

WC

,

Sheen

YY

,

Cho

MH

.

2001

.

Effects of nonylphenol, bisphenol A, and their mixture on the vivipaours swordtail fish (Xiphophorus helleri)

.

Environ Toxicol Chem

20

:

787

–

795

.

40

Brian

JV

,

Harris

CA

,

Scholze

M

,

Backhaus

T

,

Booy

P

,

Lamoree

M

,

Pojana

G

,

Jonkers

N

,

Runnalls

T

,

Bonfà

A

,

Marcomini

A

,

Sumpter

JP

.

2005

.

Accurate prediction of the response of freshwater fish to a mixture of estrogenic chemicals

.

Environ Health Perspect

113

:

721

–

728

.

41

Lindholst

C

,

Pedersen

KL

,

Pedersen

SN

.

2000

.

Estrogenic response of bispehnol A in rainbow trout (Oncorhynchus mykiss)

.

Aquat Toxicol

48

:

87

–

94

.

42

Wang

H

,

Wang

J

,

Wu

T

,

Qin

F

,

Hu

X

,

Wang

L

,

Wang

Z

.

2011

.

Molecular characterization of estrogen receptor genes in Gobiocypris rarus and their expression upon endocrine disrupting chemicals exposure in juveniles

.

Aquat Toxicol

101

:

276

–

287

.

43

Chitra

KC

,

Latchoumycandane

C

,

Mathur

PP

.

2003

.

Induction of oxidative stress by bisphenol A in the epididymal sperm of rats

.

Toxicology

185

:

119

–

27

.

44

Meeker

JD

,

Ehrlich

S

,

Toth

TL

,

Wright

DL

,

Calafat

AM

,

Trisini

AT

,

Ye

X

,

Hauser

R

.

2010

.

Semen quality and sperm DNA damage in relation to urinary bisphenol A among men from an infertility clinic

.

Reprod Toxicol

30

:

532

–

539

.

45

Hatef

A

,

Alavi

SMH

,

Butts

IAE

,

Policar

T

,

Linhart

O

.

2011

.

The mechanisms of action of mercury on sperm morphology, Adenosine‐5'‐triphosphate content and motility in Perca fluviatilis (Percidae; Teleostei)

.

Environ Toxicol Chem

30

:

905

–

914

.

46

Miura

T

,

Yamauchi

K

,

Takahashi

H

,

Nagahama

Y

.

1992

.

The role of hormones in the acquisition of sperm motility in salmonid fish

.

J Exp Zool

261

:

359

–

363

.

47

Alavi

SMH

,

Cosson

J

.

2006

.

Sperm motility in fishes: (II) Effects of ions and osmotic pressure

.

Cell Biol Int

30

:

1

–

14

.

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