Chemosphere 112 (2014) 275–281

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Assessing developmental toxicity and estrogenic activity of halogenated bisphenol A on zebrafish (Danio rerio) Maoyong Song a,⇑, Dong Liang a, Yong Liang b,c, Minjie Chen b, Fengbang Wang a, Hailin Wang a, Guibin Jiang a a b c

State Key Laboratory of Environmental Chemistry and Ecotoxicology, Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences, Beijing 100085, PR China School of Medicine, Jianghan University, Wuhan 430056, PR China Key Laboratory of Optoelectronic Chemical Materials and Devices of Ministry of Education, Jianghan University, Wuhan 430056, PR China

h i g h l i g h t s  According to LC50 values, the rank order of toxicities were TCBPA > TBBPA > BPAF.  Three H-BPAs exposure resulted in a variety of developmental lesions in the embryos/larvae.  BPAF shows a stronger estrogenic activity than BPA both in in vivo and in vitro.  TCBPA and TBBPA show no estrogenic activity both in in vivo and in vitro.

a r t i c l e

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Article history: Received 10 February 2014 Received in revised form 17 April 2014 Accepted 22 April 2014

Handling Editor: A. Gies Keywords: Halogenated bisphenol A Developmental toxicity Estrogenic activity Vitellogenin MVLN

a b s t r a c t Halogenated bisphenol A (H-BPAs), widely used in industrial production, have been identified in various environmental matrices and detected in human serum and breast milk. The persistence and prevalence of H-BPAs in the environment underscore the need to in-depth understand their adverse effects to humans and other organisms. In the present study, zebrafish embryos/larvae were used as models to investigate the developmental toxicities of three H-BPAs, namely tetrabromobisphenol A (TBBPA), tetrachlorobisphenol A (TCBPA), and bisphenol AF (BPAF). The half lethal concentration (LC50) values indicated that the rank order of toxicities of the chemicals were TCBPA > TBBPA > BPAF. Three H-BPAs exposure resulted in a variety of developmental lesions in the embryos/larvae, such as a delay in time to hatch, edema, and hemorrhage. The estrogenic activities of H-BPAs were determined by means of in vivo vitellogenin (vtg) assay and in vitro MVLN assay. Here only BPAF specifically shows a stronger estrogenic activity than BPA both in in vivo and in vitro. These data suggest that TCBPA, TBBPA, and BPAF are more potent toxicants than BPA, and indicate that further research of the mechanisms on their toxicities is required. Ó 2014 Published by Elsevier Ltd.

1. Introduction Bisphenol A (BPA) is an important industrial chemical which is widely used in the manufacture of polycarbonated plastic, food can linings, chemical papers, and dentistry sealants (Bermudez et al., 2010; Riu et al., 2011). Tetrabromobisphenol A (TBBPA) and tetrachlorobisphenol A (TCBPA) are halogenated derivatives of BPA (HBPAs), which feature bromine or chorine substituents on the phenolic rings, are used as flame retardants (Hakk and Letcher, 2003; Chu et al., 2005; de Wit et al., 2010). Bisphenol AF (BPAF) is a fluo-

⇑ Corresponding author. Tel./fax: +86 10 62849178. E-mail address: [email protected] (M. Song). http://dx.doi.org/10.1016/j.chemosphere.2014.04.084 0045-6535/Ó 2014 Published by Elsevier Ltd.

rinated derivative of BPA in which the methyl groups bound to the central bridging carbon atom are replaced by CF3 groups. BPAF is commonly used in polycarbonate copolymers in high-temperature composites, electronic materials, and gas-permeable membranes (Matsushima et al., 2010; Li et al., 2012). High production and wide applications facilitate the release of H-BPAs into environment. TBBPA has been found in river sediment (Watanabe et al., 1983; Zhang et al., 2009), sewage sludge (Sellström and Jansson, 1995; Alaee et al., 2003), air samples (Sjödin et al., 2001), and wildlife (Morris et al., 2004). It was also detected in human blood, and potential sources for human exposure are indoor environment and food (Jakobsson et al., 2002; Thomsen et al., 2005). TCBPA was detected in waste-paper recycling plants (Fukazawa et al., 2002), sediment and sludge

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(Voordeckers et al., 2002; Chu et al., 2005). Recently, it was detected in drinking water at a low concentration (Fan et al., 2013). BPAF was detected in about three-fourths of collected surface water and sewage samples in Germany, as well as in more than half of the sediment samples (Feng et al., 2012). Song et al. (2012) reported that indoor dust samples collected from the local household contained BPAF with concentrations ranging from 15.5 to 739 ng/g dw. Because of high lipophilicity, low volatility, low water solubility and bio-recalcitrance (Voordeckers et al., 2002; Hakk and Letcher, 2003), these chemicals are persistent and prevalent contaminants in the environment. Bisphenol A has been shown to induce estrogen-dependent response via binding to estrogen receptors (ERs) (Rubin, 2011), and this estrogenic activity is likely involved in the outset of many of its reverse effects in animal models (O’Connor and Chapin, 2003; Vandenberg et al., 2009). BPAF has the ability to bind with ERs, and the binding affinity or estrogenic activity was about one order or magnitude stronger than dose BPA (Kitamura et al., 2005; Okada et al., 2008; Matsushima et al., 2010). The higher estrogenic activity of BPAF than BPA was attributed to that the trifluoromethyl group is much more electronegative than the methyl group, which is related to its estrogenic activity (Kitamura et al., 2005; Matsushima et al., 2010; Bermudez et al., 2010). The estrogenic activity of TCBPA was similar to or higher than that of BPA (Meerts et al., 2001; Kitamura et al., 2005; Vandenberg et al., 2009), however, studies on the estrogenic activity of TBBPA in vitro are often inconsistent (Yamasaki et al., 2003; Bay et al., 2004; Wetherill et al., 2007; Akahori et al., 2008). Meerts et al. (2001) reported that brominated BPA analogs are not as estrogenic as BPA, and the potency of brominated-BPAs as ER agonists decreases as the number of bromine atoms increases. The estrogen-dependent production of the yolk precursor vitellogenin (vtg) was not affected by exposure to TBBPA in rainbow trout and eelpout (Christiansen et al., 2000; Ronisz et al., 2004), but its exposure effects including reduction of egg production, survival and overall reproductive success were observed in zebrafish (Kuiper et al., 2007). Because of their widespread use and presence in environmental samples, H-BPAs are currently under study to determine whether they also cause negative effects in wildlife and human systems. However, information on their toxicological outcomes, environmental presence and environmental fate are still limited. Zebrafish (Danio rerio) is an excellent vertebrate model for assessing the toxicity of novel compounds, pollutants and pharmaceuticals (Reimers et al., 2004; Ali et al., 2011). The developing embryo and larvae are generally considered to be the most sensitive stage in the life cycle of zebrafish, being particularly sensitive to low-level environmental pollutant (Schulte and Nagel, 1994). In this study, zebrafish embryos and adults were employed as models for evaluating the toxicity of H-BPAs, with particular emphasis on their estrogenic activity. The estrogenic potential of four chemicals, including BPA, TBBPA, TCBPA, and BPAF, was determined via in vivo vtg assays in male zebrafish and in vitro MVLN assays.

2. Materials and methods 2.1. Chemicals TCBPA and BPAF were purchased from TCI (Portland, OR). 17bestradial (E2), BPA and TBBPA were purchased from Sigma–Aldrich (USA). All chemicals have a purity of 98% or greater unless otherwise mentioned. We dissolved BPA, TCBPA, TBBPA and BPAF in dimethyl sulfoxide (DMSO) to form stock solutions and stored them away from light. All other chemicals used in this study were analytical grade.

2.2. Treatment of zebrafish adult/embryos Zebrafish embryos were collected as previously described (Jin et al., 2009; Li et al., 2011; Chan and Chan, 2012) and rinsed with an embryo medium (0.137 M NaCl, 5.4 mM KCl, 0.25 mM Na2HPO4, 0.44 mM KH2PO4, 1.3 mM CaCl2, 1.0 mM MgSO4 and 4.2 mM NaHCO3). Healthy embryos were selected with a dissecting microscope and incubated in dishes at 28 ± 0.5 °C until chemical treatment. The embryos were exposed in 4 mL glass vials to BPA, BPAF, TCBPA, and TBBPA at different concentrations according to the preliminary study. Chemicals concentrations in exposure water were analyzed by liquid chromatography/mass spectrometry (LC/ MS) (Song et al., 2014). A vehicle and a nonvehicle control were included to confirm that DMSO concentration used had no effect on zebrafish development. Studies were conducted with 30 embryos per dose and all studies were repeated at least three times. Developmental lesions, including edema and hemorrhage, as well as death and date of hatching, were recorded daily. Two-month-old male zebrafish was cultured in recirculating aquarium tanks at 28 ± 0.5 °C. The fish was maintained in a 14:10 light/dark cycle and fed twice daily with fresh Artemia nauplii. Exposure studies were performed in 10 L glass aquaria. Ten males selected randomly from acclimatized fish were put into each aquaria and triplicate aquaria were used for each exposure concentration. The stock solutions were added into the exposure tanks using a dilution apparatus. Static conditions were applied for the exposures and the total volume of solution was changed every day. The exposure duration was 21 d and the fish were fed daily.

2.3. Measurement of vtg in zebrafish blood The blood of zebrafish was collected in freshly heparinized microhematocrit tubes by severing the caudal peduncle, and centrifuged at 11 000 rpm for 10 min. After centrifugation, serum volumes (2–10 lL) were obtained from individual adult zebrafish, and subsequently diluted in phosphate-buffered saline (PBS) and frozen at 80 °C until analysis. Vtg was purified from serum of E2induced zebrafish according our previous work (Shao et al., 2005; Song et al., 2005), and used as a standard after identification and quantification. Serum vtg levels were determined using a comparative enzyme-linked immunosorbent assay (ELISA) (Shao et al., 2005).

2.4. Estrogenic activity analysis in vitro The relative estrogenic activity is determined by MVLN assay. This cell line was stably transfected with the luciferase reporter gene and estrogen-responsive element, and ER agonists can induce the production of luciferase. Estrogenic activity of each sample was determined as previously described (Song et al., 2006). Briefly, 60 interior wells of a 96-well culture plate were each seeded with 250 lL cell suspension at a density of about 5  104 cells per well. Cells were cultured under aseptic conditions in a humidified CO2 incubator at 37 °C and 5% CO2. The cells were starved in steroidfree medium for 48 h before exposure. A concentration range of 0.1 pM to 10 nM E2 was used as a positive control. Luciferase activity was measured using the LucLite kit according to the manufacturer’s protocol. Luminescence was read by Microplate Reader (Varioskan Flash) after cells were exposed to test compound for 48 h, and total protein content was simultaneously measured by the Bradford assay to normalize luminescent units. The maximal luciferase activity induction of E2 was set as 100%, and the responses of other samples were converted to a percentage of the maximum level. Four replicates were used in each experiment.

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2.5. Statistical analysis SPSS statistical software (Version 13.0) and Sigma Plot 10.0 were used for statistical analysis. The significant differences between control and treated groups were determined using a one-way analysis of variance and Tukey’s multiple range test. Differences were statistically significant if p < 0.05. 3. Results 3.1. Relative toxicity of BPA and H-BPAs to zebrafish embryo development The time to hatch was measured in developing zebrafish embryos exposed to BPA, BPAF, TCBPA or TBBPA and compared to control embryos. The results of LC/MS indicated that these chemicals were well-distributed in the aqueous treatment. The embryos of control group developed normally in embryo medium, and almost all embryos were hatched between 48 and 72 h (data not shown). As shown in Fig. 1, there were no difference in hatching rate and hatching time of embryos exposed to BPA at three exposure concentration (0.5, 1.0, and 1.5 mg/L). The exposure of BPAF (1.0 and 1.5 mg/L) and TBBPA (0.5 and 1.0 mg/L) could delay the occurrence of hatching. The hatching rates of embryos were significantly inhibited by higher concentration (1.5 mg/L) of TCBPA and TBBPA at 72 hpf, indicating a potential developmental toxicity. The impact of BPA, BPAF, TCBPA or TBBPA exposure on embryos/larvae survival was assessed. More than 90% embryos/larvae died when treated with 10 mg/L of BPA for 96 h (Fig. 2a). However, all embryos/larvae exposed to 5 mg/L or lower concentrations of BPA survived for 144 h. All embryos/larvae died when exposed to 2 mg/L BPAF for 144 h, and lower exposure concentrations had no impact on embryos/larvae mortality (Fig. 2b). TCBPA exposure

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was more acutely toxic than BPA and BPAF, resulting in 100% mortality of 1.0 mg/L exposed embryos/larvae by 120 hpf, and the 1.5 mg/L dose by 96 hpf (Fig. 2c). TBBPA was also more toxic than BPA and BPAF, resulting in 100% mortality in embryos/larvae exposed to 1.5 mg/L dose by 144 hpf (Fig. 2d). Compared with hatching results, we considered that TCBPA and TBBPA at 1.5 mg/L dose were more toxic in both embryonic zebrafish and juvenile period of zebrafish development, while TCBPA at 1.0 mg/L was more toxic than other chemicals to zebrafish larvae. In order to rank these compounds in order of toxicity, the half lethal concentration (LC50) values for BPA, BPAF, TCBPA, and TBBPA exposure were determined by the percent mortality. The 144 h LC50 values of BPA, BPAF, TCBPA, and TBBPA for zebrafish larvae were 7.5, 1.75, 0.75 and 1.24 mg/L, respectively. From the most toxic to the least toxic, the rank order of toxicities of the chemicals were TCBPA > TBBPA > BPAF > BPA. The lowest observed adverse effect level (LOAEL) was used to indicate levels of exposure at which development was affected. The LOAELs were 5.0 mg/L for BPA while 1.0 mg/L for BPAF and TCBPA, and 0.5 mg/L for TBBPA. These results illustrate that H-BPAs are more toxic than BPA. 3.2. Lesion occurrence in exposure embryos/larvae Exposure to BPA and H-BPAs resulted in developmental lesions in zebrafish embryos/larvae. As shown in Table 1, lesions, such as yolk sac (YS) and pericardial (PC) edema, and hemorrhage detected at 0.5, 1.0 and 1.5 mg/L of BPA exposure did not differ from control. PC edema could be observed when BPA exposure dose was higher than 10 mg/L (data not shown). BPAF (1.5 mg/L) exposure resulted in significant PC edema only, while TBBPA (1.5 mg/L) resulted in both significant YS and PC edema. Additionally, zebrafish embryos/larvae exposed to TBBPA (1.0 and 1.5 mg/L) had an increase in hemorrhage as compared to control. Significant YS

Fig. 1. The hatching rates in zebrafish embryos exposed to BPA, BPAF, TCBPA, and TBBPA at various concentrations.

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Fig. 2. Mortality of zebrafish embryos/larvae exposed to BPA, BPAF, TCBPA, and TBBPA at various concentrations.

Table 1 Development lesions (%) observed in zebrafish embryos/larvae exposed to BPA, BPAF, TCBPA, and TBBPA. DMSO

YS edema PC edema Hemorrhage

4.4 2.2 0

BPA (mg/L)

BPAF (mg/L)

TCBPA (mg/L)

TBBPA (mg/L)

0.5

1.0

1.5

0.5

1.0

1.5

0.5

1.0

1.5

0.5

1.0

1.5

0 2.2 0

5.6 0 2.2

4.4 5.6 0

3.3 0 0

5.6 3.3 0

4.4 37.8 1.1

2.2 0 0

36.7 41.1 34.4

65.6 33.3 55.6

4.4 0 2.2

6.2 4.4 12.2

26.7 45.6 42.2

Lesions were record and present as lesion percentage occurring at 72 hpf. All data are representative of three replicate experiments (n = 30 per exposure dose).

and PC edema, and hemorrhage were observed at 1.0 and 1.5 mg/L dose of TCBPA exposure. Bent and hook-like tails were not observed in all embryos/larvae treated with all test concentrations of four chemicals in 144 hpf. These data show that exposure to BPAF, TCBPA, or TBBPA results in development lesions at a lower dose than BPA exposure. 3.3. Estrogenic activity of BPA and H-BPAs In order to investigate the estrogenic activity of these chemicals, the vtg level in adult male zebrafish was measured using a semiquantitative ELISA method. During 21 d of exposure to four chemicals, no increase in mortality was observed for BPA, BPAF and TBBPA compared to the control at all test conditions (Fig. 3a). A significant mortality (30%, p < 0.05) was observed at the 1.5 mg/L TCBPA exposure (Fig. 3a), indicating that it was more toxic to adult zebrafish than other three chemicals. The results of the vtg analysis in male zebrafish following 21 d exposure period to BPA, BPAF, TCBPA, and TBBPA are given in Fig. 3b. The hormone

E2 was used as positive control (PC) in studies of the estrogenic effects of chemicals on the induction of vtg. The male zebrafish was highly responsive towards E2 treatment and a significant (p < 0.001) vtg induction was observed at a concentration of 5 lg/L. BPA and BPAF induced plasma vtg in a concentrationdependent manner. Vtg levels in the groups of 0.5, 1.0, and 1.5 mg/L BPA were 7.3 ± 0.7, 38.4 ± 1.1, and 696.6 ± 52.69 lg/mL, respectively. Due to the relatively high estrogenic activity of BPAF, vtg levels in the groups of 0.5, 1.0, and 1.5 mg/L were 932.4 ± 24.8, 2236.2 ± 149.5, and 5292.3 ± 248.6 lg/mL, respectively. No significant vtg induction above negative control (NC, DMSO control) fish was observed in all groups treated with TCBPA and TBBPA. The estrogenic activities of BPA and H-BPAs were also determined using an estrogen receptor-mediated, chemical-activated luciferase reporter gene-expression assay. As shown in Fig. 4, the induction of luciferase activity by TCBPA and TBBPA were at the basal level of control, whereas BPA and BPAF induced luciferase activity in a dose-related manner. BPAF showed higher estrogenic activity than did BPA in MVLN assay, with maximum induction

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Fig. 3. (a) Survival of zebrafish after 21 d exposure to BPA and H-BPAs. Data are given as average ± SD of two replicate aquaria. (b) Vtg analyzes in male zebrafish after 21 d exposure to BPA and H-BPAs (average with SD; n = 10). NC: negative control (exposure to 0.5% DMSO); PC: positive control (exposure to 5 lg/L E2). Statistical evaluation by ANOVA: *p < 0.05, **p < 0.001 compared to NC.

Fig. 4. Dose-repose luciferase activities of BPA and H-BPAs in the MVLN assay. The relative estrogenic activities are expressed as mean ± SD of triplicate measurements in one representative experiment. * Significant cytotoxicity was observed at test concentration.

values of 72.3% at 5 lM and 59.4% at 40 lM, respectively. In the present study, the estrogenic activities of BPA and H-BPAs revealed by the MVLN assay were in accordance with the results of vtg assay in zebrafish blood. 4. Discussion As widely used raw materials in manufacture, BPA and H-BPAs may leak into the environment and possibly accumulate in biological systems. The halogenated derivatives of BPA have been found

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in soil, water, air, and human plasma (Oberg et al., 2002; Thomsen et al., 2002; Fukazawa et al., 2002; Labadie et al., 2010; Song et al., 2012). In our recent work, the concentrations of H-BPAs in sewage sludge were 259, 142.5 and 45.1 ng/g dw for TBBPA, TCBPA and BPAF (Song et al., 2014). We also found that the concentrations of TBBPA and TCBPA in river water were 143 ± 3.7 and 224 ± 11 ng/L, respectively (Yin et al., 2011). The exposure concentrations of H-BPAs used in this study were much higher than their environmental realistic concentrations. However, these chemicals were persistent and prevalent contaminant in the aquatic environment (Voordeckers et al., 2002; Hakk and Letcher, 2003), and limited information concerning the toxicological impacts of these chemicals is available. Therefore, our results could be used as a parameter in the evaluation of water quality. Our data show an increase in embryos mortality following developmental exposure to H-BPAs. H-BPAs proved to be 5–10 times more potent than BPA, resulting in 100% mortality, indicating that H-BPAs were more acutely toxic to zebrafish embryos than BPA. BPA and TBBPA were reported toxic to aquatic organisms, but few studies were performed to show effects of exposure of TCBPA and BPAF. Here we showed that TCBPA is a more potent chemical than either TBBPA or BPAF. Our results of post-hatch survival in exposed embryos/larvae are in accord with the findings from previous work examining post-hatch survival of embryos at 3 lM (1.6 mg/L) of TBBPA with 100% mortality (McCormick et al., 2010). We found that H-BPAs exposure results in a variety of developmental lesions in the embryos/larvae, such as a delay in time to hatch, and vascular lesions (edema and hemorrhage). Although recent work reported that embryos exposed to BPA and TBBPA would lead to tail malformation (Duan et al., 2008; McCormick et al., 2010), it was not observed in all embryos/larvae in this study. There is sufficient qualitative information on reproductive and developmental toxicity of BPA to aquatic organisms (Crain et al., 2007). All of these effects of BPA have been attributed to effects on steroid hormone receptors such as ER (Matsushima et al., 2010). BPA and BPAF have been shown to induce estrogen-dependent response in vitro via binding to ER (Song et al., 2006; Bermudez et al., 2010), whereas studies on the estrogenic activity of TBBPA in vitro are inconsistent. The estrogenic activity of TBBPA was examined by several investigators using various in vitro methods. While some observed that TBBPA exhibited weak ER effect in vitro (Li et al., 2010), others found that TBBPA did not exert any effects, even at high concentrations (Dorosh et al., 2011; Lee et al., 2012). The MVLN assay has been widely used to study ER activity of chemicals (Freyberger and Schmuck, 2005). In this study, we used the MVLN assay to investigated ER activity and estrogenic potency of H-BPAs. BPA exhibits weak estrogenic activity in the assays, and BPAF exerts a higher estrogenic potency than BPA, indicating that BPAF binds to ER more strongly than does BPA. It is consistent with the results of estrogenic activity in vitro for BPA and BPAF reported in previous work (Kitamura et al., 2005; Okada et al., 2008). The potential estrogenic activity of BPAF is of concern in part because its CF3 group is much more electronegative (and potentially reactive) than the CH3 group of BPA (Matsushima et al., 2010). Compared with BPA and BPAF, both TCBPA and TBBPA showed no estrogenic activity. It is likely that bromine or chorine substituents on the phenolic rings reduce the binding activity of TCBPA and TBBPA for ER. However, the underlying mechanisms for the estrogenic activity of H-BPAs with different molecular structure have not been completely clarified. The vtg assay is a frequently used in vivo biomarker for estrogenic activity in oviparous vertebrates. In the present study, we further examined the estrogenic activity of BPA and H-BPAs in vivo by means of vtg assay. It has been reported that BPA induces synthesis of vtg in many fish species (Tabata et al., 2004; Brian et al., 2005). Here alteration of vtg confirms the potential for BPA

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and BPAF to exert estrogenic effects on male zebrafish. Chow et al. recently reported that TBBPA gave a 60-fold induction of vtg1 mRNA levels, which suggested that TBBPA was more estrogenic to embryos compared with BPA (Chow et al., 2013). However, there was no vtg induction in rainbow trout and eelpout by intraperitoneal exposure of TBBPA (Christiansen et al., 2000; Ronisz et al., 2004). Similar equivocal results were also obtained in other in vivo studies. Kitamura et al. (2005) exposed TBBPA to ovariectomized B6C3F1 mice by intraperitoneal injection and noted increased uterus to body weight ratio in all exposed groups, suggesting a weak estrogenic activity of TBBPA. However, an uterotrophic assay conducted by Ohta et al. (2012) showed that TBBPA was negative for estrogenic responses by both routes of oral gavage and subcutaneous injection using concentrations up to 1000 mg/ kg bw/d. In the present study there was no indication for an increase of production of vtg in zebrafish blood after 21 d exposure of TBBPA. Similarly, no vtg induction was observed for TCBPA. In accordance with in vitro testing, the present assay identified BPAF as relatively strong xenoestrogens. Further, when observing the chemical structure of the TCBPA and TBBPA, the negative response appeared plausible. The estrogenic activity of TCBPA and TBBPA is likely not playing a role in their developmental toxicities. However, it was found that both TCBPA and TBBPA exhibited significant thyroid hormonal activity towards rat pituitary cell line GH3, which releases growth hormone in a thyroid hormone-dependent manner (Kitamura et al., 2005; Sun et al., 2009). Since thyroid hormones influence reproduction in aquatic vertebrates and may interact during larval development (Arcand-Hoy and Benson, 1998), the presence of TCBPA and TBBPA in aquatic environment will impact on Zebrafish embryo development due to their thyroid hormonal activity. Moreover, some in vitro studies of TBBPA have shown immunological effects (Darnerud, 2003), and inhibition of synaptic neurotransmitter uptake (Mariussen and Fonnum, 2003). Recently, TBBPA has specifically been shown to inhibit sarcoplasmic/endoplasmic reticulum Ca2+-ATPases at low concentrations and also active the Ryanodine receptor Ca2+ channel, which may be the underlying of its cytotoxicity (Ogunbayo and Michelangeli, 2007; Ogunbayo et al., 2008). 5. Conclusion The data present here demonstrate that developmental exposure to H-BPAs result in a reduction in embryos/larvae survival and an increase occurrence in a variety of developmental lesions, whereas BPA exposure at the same concentration does not appear to be developmental toxic. There are differences in their potency, with TCBPA and TBBPA being more acutely toxic than BPAF. However, both the in vitro MVLN assay and in vivo vtg assay used here showed that BPAF specifically showed a stronger estrogenic activity than BPA, whereas TCBPA and TBBPA did not show any estrogenic activity. The information obtained in this study will be very useful in determining the water quality guidelines for these test chemicals in river and lakes, however, the mechanisms of their toxicities need further understand. Acknowledgments This work supported by the National Natural Science Foundation of China (Nos. 21377146 and 21125523), and the grants from the National Basic Research Program of China (Nos. 10CB933502 and 11CB936001). References Akahori, Y., Nakai, M., Yamasaki, K., Takatsuki, M., Shimohigashi, Y., Ohtaki, M., 2008. Relationship between the results of in vitro receptor binding assay to

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