Aquatic Toxicology 164 (2015) 145–154

Contents lists available at ScienceDirect

Aquatic Toxicology journal homepage: www.elsevier.com/locate/aquatox

Acute and chronic toxicity of tetrabromobisphenol A to three aquatic species under different pH conditions Qun He, Xinghao Wang, Ping Sun ∗ , Zunyao Wang ∗ , Liansheng Wang State Key Laboratory of Pollution Control and Resources Reuse, School of the Environment, Nanjing University, Nanjing 210023, PR China

a r t i c l e

i n f o

Article history: Received 24 February 2015 Received in revised form 3 May 2015 Accepted 4 May 2015 Available online 6 May 2015 Keywords: Tetrabromobisphenol A Aquatic organisms pH Acute toxicity Oxidative stress

a b s t r a c t Tetrabromobisphenol A (TBBPA) is a well-known brominated flame retardant. It has been detected in the environment and shows high acute toxicity to different organisms at high concentrations. In this work, the effects of pH and dimethyl sulfoxide (DMSO) on the acute toxicity of TBBPA to Daphnia magna and Limnodrilus hoffmeisteri were tested, and the oxidative stress induced by TBBPA in livers of Carassius auratus was assessed using four biomarkers. The integrated biomarker response (IBR) was applied to assess the overall antioxidant status in fish livers. Moreover, fish tissues (gills and livers) were also studied histologically. The results showed that low pH and DMSO enhanced the toxicity of TBBPA. Furthermore, changes in the activity of antioxidant enzymes and glutathione level suggested that TBBPA generates oxidative stress in fish livers. The IBR index revealed that fish exposed to 3 mg/L TBBPA experienced more serious oxidative stress than exposed to acidic or alkaline conditions. The histopathological analysis revealed lesions caused by TBBPA. This study provides valuable toxicological information of TBBPA and will facilitate a deeper understanding on its potential toxicity in realistic aquatic environments. © 2015 Elsevier B.V. All rights reserved.

1. Introduction Tetrabromobisphenol A (TBBPA) is widely used as a reactive or additive brominated flame retardant (BFR) and is incorporated into polymers to enhance its flame retardant ability. Of all BFRs, TBBPA has been produced at the highest volume. The estimated annual market demand for TBBPA from 2001 to 2003 was >200,000 ton year−1 (de Wit et al., 2010). Generally, some of the reactive flame retardants may not be chemically polymerized and can be released into the environment (de Wit, 2002). Although, the TBBPA concentration in water is still extreme lack of data, there are a large number of data showing high TBBPA concentrations in the sediments (Yang et al., 2013). For example, 0.0006 ng/g dry weight (d.w.) in sediments collected from Arctic (de Wit et al., 2010), 0.5–140 ␮g/kg d.w. in river sediments in Osaka, Japan (Watanabe et al., 1983), 270 ␮g/kg d.w. in sediments from a river downstream near a plastics production facility in Sweden (Sellstrom and Jansson, 1995), and 9800 ␮g/kg d.w. in freshwater sediments acquired from the River Skerne in northeast England (Morris et al., 2004).

∗ Corresponding authors at: State Key Laboratory of Pollution Control and Resources Reuse, School of the Environment, Xianlin Campus, Nanjing University, Nanjing, Jiangsu 210023, PR China. Tel.: +86 25 89680358; fax: +86 25 89680358. E-mail addresses: [email protected] (P. Sun), [email protected] (Z. Wang). http://dx.doi.org/10.1016/j.aquatox.2015.05.005 0166-445X/© 2015 Elsevier B.V. All rights reserved.

Nowadays, there is a hot debate about the health consequences of exposures to TBBPA (Decherf and Demeneix, 2011). Some researchers reported that TBBPA posed some latent risks to multiple aquatic species (Hu et al., 2009a; Kuiper et al., 2007b). For example, it has been claimed to have high acute toxicity to aquatic organisms, such as algae, mollusks, crustaceans, and fish (Reistad et al., 2006; Ronisz et al., 2004). Moreover, Yang et al. (2012) presented the LC50 values of TBBPA for eight resident aquatic organisms in China, and Hu et al. (2015) found that scallop Chlamys farreri after TBBPA exposure underwent significant inhibition on microsomal cytochrome P450 and b5 levels in gills and digestive gland. Chronic exposure to TBBPA in mammals could have physiological effects and induce disease, especially in relation to its thyroid-disrupting toxicity (Yang et al., 2013). However, according to some other risk assessment reports, TBBPA is not considered as a persistent and bioaccumulative toxicant (ECB, 2006, 2008; Colnot et al., 2014). More, fast absorption from gastrointestinal tract, rapid metabolism and no accumulation in lipid tissues have been observed in toxicokinetic studies with mammals (Schauer et al., 2006; Kuester et al., 2007). TBBPA is lipophilic (logKow = 4.50) and has low water solubility; however, it is very soluble in methanol, acetone, and dimethyl sulfoxide (DMSO) (Birnbaum and Staskal, 2003). Two phenolic hydroxyl groups in TBBPA are important for increasing its water ∗ ) at higher pH. The S ∗ of TBBPA at pH 9.50 is solubility (Sw w

146

Q. He et al. / Aquatic Toxicology 164 (2015) 145–154

27,900 ± 400 mg/L, which is higher than that at pH 3.05 by five orders of magnitude (Kuramochi et al., 2008; Osako et al., 2004). Studies on TBBPA in aquatic systems have mostly been performed at pH near the pKa2 of TBBPA (8.5), and the behavior of TBBPA in acidic water may be different for that near pH 8.5 (WHO/ICPS, 1995). Thus, it is necessary to investigate the potential toxicity of TBBPA to different aquatic organisms under different pH conditions. Due to the low solubility of TBBPA, almost all the toxicological studies of TBBPA have been conducted with co-solvents (Carlsson and Norrgren, 2014; Hamers et al., 2006; van Boxtel et al., 2008). However, in the natural environment, it is not common for TBBPA to coexist with co-solvents; therefore, it is necessary to compare the toxicological effects of TBBPA in the presence or absence of co-solvents under different pH conditions. Detoxification enzymes can serve as indicators of the level of environmental pollution because of their close correlation with pollutant concentrations (Yang et al., 2013). A panel of multiple biomarkers in fish can provide a general assessment of oxidative stress and is a useful tool for developing more comprehensive ecotoxicological profiles of chemicals. In addition, the effects of chemicals on biomarkers in fish can help to explain the physiological processes that are disrupted by those chemicals and also help to discover possible mechanisms of detoxification. Thus, it is feasible to discuss the potential toxic effects of TBBPA on fish livers using multiple biomarkers. In the present study, to analyze the influence of pH on the toxicity of TBBPA and toxic effects of TBBPA on different species, different pH-regimes and three major taxonomic groups were chosen to conducted experiments. The acute toxicity of TBBPA and TBBPA dissolved in DMSO to Daphnia magna (D. magna) and Limnodrilus hoffmeisteri (L. hoffmeisteri) under different pH conditions was measured to confirm whether there is synergistic effect between TBBPA and DMSO. Furthermore, four oxidation-related biomarkers, the activity of superoxide dismutase (SOD), catalase (CAT), and glutathione S-transferase (GST) as well as the level of glutathione (GSH) were measured in fish livers under different pH conditions. Integrated biomarker response (IBR) indices were calculated to estimate the integral effects of TBBPA-induced oxidative stress under different pH conditions in fish livers. Finally, to better account for the toxicological impacts of TBBPA, liver and gill tissues were examined histologically. 2. Materials and methods 2.1. Test chemicals DMSO, hydrochloric acid, and sodium hydroxide were of analytical grade and were obtained from Sinopharm Chemical Reagent Co., Ltd. (Shanghai, China). TBBPA and 3-(N-morpholino) propanesulfonic acid (MOPS) a purity of 98% and 99%, respectively, were both supplied from Aladdin® Reagent (Shanghai, China). The kits for the analysis of the oxidative stress biomarkers were purchased from Nanjing Jiancheng Bioengineering Institute (Nanjing, China). 2.2. Test organisms Organisms of three aquatic species that serve as representatives of three major taxonomic groups (Crustaceans, Zoobenthos,

and Vertebrates) were chosen as the test organisms for this study. The D. magna strain was supplied by the Research Center for EcoEnvironmental Sciences, Chinese Academy of Sciences (Beijing, China). Tap water filtered by activated carbon and aerated for more than 48 h was used as the culture medium. Daphnia organisms were maintained in the culture water (pH 7.50 ± 0.25) under a light–dark cycle of 16:8 h at 20 ± 1 ◦ C and were fed three times a day with green algae, Scenedesmus obliquus. The culture medium was renewed three times a week. The experiments were conducted with juvenile fleas (6–24 h old) that had undergone three generations of parthenogenesis. To guarantee the sensitivity of D. magna to toxins, acute toxicity tests were performed using K2 Cr2 O7 as the reference toxicant before performing the experiments (OECD, 2004). L. hoffmeisteri organisms (length: 5–6 cm) that were slim and red–brown in color were obtained from an aquatic products market (Jinan, China) and cultured in a large aquarium containing some river sediments with a 16:8 h light–dark cycle at 23 ± 1 ◦ C. After a 7 day acclimation period, L. hoffmeisteri organisms that were intact and of similar length and size were chosen for the subsequent toxicity tests. Carassius auratus (weight: 30.15 ± 0.5 g; length: 13.87 ± 2 cm) was purchased from a local aquatic breeding center. Before the experiments, the goldfish was acclimatized for 10 days in tanks containing 150 L dechlorinated and aerated water at 20 ± 1 ◦ C. The fish were fed with commercial pellets, and food residue and feces were removed twice a day. 2.3. Water quality The water quality parameters of the tap water used for the acclimation period and the subsequent experiments were measured and are listed in Table 1.

2.4. Toxicity of TBBPA to D. magna and L. hoffmeisteri The tests for the acute toxicity of TBBPA to D. magna were carried out according to the National Standard Method of China (water quality-determination of the acute toxicity of substances to Daphnia (D. magna straus) GB/T 13266-1991). Duration of the tests was 24 h with immobilization as the endpoint. There were three replications for per treatment and ten animals in each treatment. Preliminary experiments were performed to investigate the effects of different pH stabilized with 3.58 mM buffering agent (MOPS) (de Schamphelaere et al., 2004) on D. magna and L. hoffmeisteri. Results showed that the activities of the organisms were not changed after 24 h of exposure to the test media at pH 6.0, 7.0, 7.5, 8.0, or 9.0, indicating that the effects of pH in this range are negligible and that MOPS was suitable for pH buffering. Based on the preliminary experiments, Scheme 1 was designed and used for the subsequent experiments. Different TBBPA concentration gradients for the different pH-regimes were decided by the preliminary experiments. All media were adjusted to be within ±0.1 of the desired. The experiments for the acute toxicity of TBBPA to L. hoffmeisteri were also conducted following Scheme 1, except that the TBBPA concentration gradients used were different (Supporting information, Table S2). Every species respond to TBBPA differently, so the concentra-

Table 1 Water quality parameters of the water used for acclimation and subsequent experiments. Water quality parameters pH Conductivity Total hardness Alkalinity Na+

7.25 ± 0.25 340.6 ± 16.4 ␮s/cm 135.5 ± 9.3 mg CaCO3 /L 40.7 ± 5.2 mg CaCO3 /L 11.2 ± 0.2 mg/L

K+ Mg2+ Ca2+ Cl− DO (dissolved oxygen)

2.34 ± 0.07 mg/L 7.74 ± 0.02 mg/L 41.07 ± 0.82 mg/L 28.3 ± 1.2 mg/L 6.76 ± 0.84 mg O2 /L

Q. He et al. / Aquatic Toxicology 164 (2015) 145–154

147

Scheme 1. The testing procedures for the acute toxicity of TBBPA to D. magna.

tion gradients for different species were also determined by their preliminary experiments. The procedures for testing the acute toxicity of TBBPA dissolved in DMSO (TBBPA-D) to D. magna and L. hoffmeisteri were also similar to Scheme 1, except that dechlorinated water at pH 6.0, 7.0, 7.5, 8.0, and 9.0 was used to dilute the corresponding TBBPA-D stock solutions to the proper concentration gradients (Supporting information, Tables S3 and S4). The percentage of DMSO in each TBBPA-D solution was lower than 0.1% (v/v). Preliminary experiments were conducted to confirm the effects of 0.1% (v/v) DMSO on the organisms and results showed no effect on the activities of the organisms. 2.5. The chronic toxicity of TBBPA to C. auratus under different pH conditions Three representative pH 6.0, 7.5 (the pH of the tap water in our laboratory), and 9.0, were selected to study antioxidant responses in fish livers. Sun et al., (2008) stated that the maximum sorption amount of TBBPA equals to the amount of desorption in water–sediment interface, at 25 ◦ C, TBBPA concentration in the range of 0.1–1.5 mg/L. So far, many literatures reported the TBBPA concentrations in sediments have been over the maximum theoretical sorption amount (Yang et al., 2013). That is, it is possible that TBBPA can be released from sediments to water, which may result in the concentrations of TBBPA to reach the range (0.1–1.5 mg/L). Besides, Yang et al. (2013) studied multibiomaker responses upon exposure to TBBPA in C. auratus at 0.35–2.00 mg/L for 64 days and Kuiper et al. (2007a) studied chronic toxicity of TBBPA to European flounder (Platichthys flesus) at 0.40 mg/L for 150 days. Therefore, 0.3 and 3 mg/L (i.e., 0.55 and 5.50 ␮M) of TBBPA were chosen as the exposure concentrations. The exposure solutions were sampled at different time points to determine the actual exposure concentra-

tions using the API 4000 instrument (AB Sciex, Framingham, USA). The detailed procedures of the LC–MS/MS analysis of TBBPA in the different exposure solutions are provided in Supporting information, and the results are presented in Table 2. Procedures for testing the chronic toxicity of TBBPA to C. auratus under different pH conditions are plotted in Scheme 2. 2.6. Biochemical assays Four biomarkers, including the activity of the antioxidant enzymes SOD, CAT, and GST as well as GSH level, were analyzed. Antioxidant enzyme activity represents the concentrations of the enzymes. SOD activity was measured at 550 nm using xanthine oxidase based on the inhibition of cytochrome c caused by the superoxide radical (McCord and Fridovich, 1969). CAT activity was determined by monitoring residual H2 O2 absorbance at 405 nm following the method of Góth (1991). GSH level was evaluated at 420 nm following the procedure of Jollow et al. (1974) by using 5,5 dithiobis-2-nitrobenzoic acid (DTNB) reagent. DTNB was reduced by free sulfhydryl groups of GSH to form yellow 5-thio-nitrobenzoic acid (TNB). GST activity was quantified at 412 nm using the Diagnostic Reagent Kits according to the manufacturer’s instructions. Protein levels were determined at 595 nm by the Coomassie Brilliant Blue dye using the Bradford method (1976). 2.7. Integrated biomarker response (IBR) IBR is a useful tool for assessing ecological risk (Damiens et al., 2007; Kim et al., 2010), and it combines all measured biomarkers to from one general stress index. The detailed calculations for the IBR analysis were based on Kim et al. (2010). (1) Standardization of data by the formula Y = (X − m)/s, where Y is the standardized value of biomarker, X is the value of a biomarker

Table 2 The nominal and measured concentrations of TBBPA in exposure solutions (mg/L).

Nominal Mean ± SD

pH(6.0)–TBBPA(0.3)

pH(6.0)–TBBPA(3)

pH(7.5)–TBBPA(0.3)

pH(7.5)–TBBPA (3)

pH(9.0)–TBBPA (0.3)

pH(9.0)–TBBPA(0.3)

0.30 0.28 ± 0.03

3.00 2.7 ± 0.47

0.30 0.28 ± 0.05

3.00 2.78 ± 0.40

0.30 0.29 ± 0.03

3.00 2.76 ± 0.32

148

Q. He et al. / Aquatic Toxicology 164 (2015) 145–154

Scheme 2. The testing procedures for oxidative stress evaluation of TBBPA in C. auratus livers.

from each treatment, m is the mean of the biomarker, and s is the standard deviation of the biomarker. (2) Using standardized data, Z was computed as Z = Y in the case of activation or Z = −Y in the case of inhibition. Then, the minimum value for each biomarker was obtained and added to Z. (3) The score (S) was computed as S = Z + |Min|, where S ≥ 0 and |Min| is the absolute value of Min. (4) A star plot radius coordinate represents the score of a given biomarker. When the Si and Si+1 are assigned as two consecutive clockwise score of a given star plot, n is assigned as the number of radii corresponding to the biomarkers. Star plot areas (Ai ) was calculated as:

Ai =

Si sinˇ(Si cosˇ + Si+1 sin␤) 2

(1)

where ␤ = Arctan(Si sin˛ − Si cos˛); ˛ = 2/n; Sn+1 = S1 . S i

(5) Summing up all values, and the corresponding IBR value is defined as IBR =

n  i=1

Ai .

2.8. Histopathological study of C. auratus gills and livers Samples of histological were soaked in 10% formaldehyde once it was being dissected. Then, samples were embedded in paraffin blocks, sliced into 5 ␮m in thickness (KD-2258 Rotary Microtome, Jinhua, China), and placed onto glass slides. After stained by hematoxylin and eosin (H&E), the slides were observed and photoed by an optical microscope (TI-U Nikon, Japan) with 40× magnification. 2.9. Statistics analyses Statistical analyses were conducted using the SPSS statistical software (PASW Statistics 18 version, Chicago, USA). The normality of the data distribution was tested by the Shapiro-Wilk, and the homogeneity of variance was determined by the Levene test. Acute toxic values were expressed as EC50 values (concentration for 50% of maximal effect, mg/L) with 95% confidence intervals and were calculated by Probit analysis. One-way analysis of variance (ANOVA) with the Duncan’s test was used to analyze differences between groups at the same time points, and the significance level was set at p < 0.05. Results are expressed as the means ± standard deviation (SD).

Q. He et al. / Aquatic Toxicology 164 (2015) 145–154

149

Table 3 The EC50 values of the toxicity of TBBPA and TBBPA dissolved by DMSO to D. magna with confidence limits (95% probability), n = 3. pH

6.0

7.0

7.5

8.0

9.0

TBBPA TBBPA + DMSO

0.45 (0.36–0.61) 0.31 (0.18–0.39)

3.97 (2.93–5.02) 0.817 (0.24–1.21)

11.54 (9.82–13.51) 2.70 (1.08–3.74)

17.34 (13.27–21.25) 8.63 (5.25–10.72)

25.33 (22.90–28.74) 18.29 (14.89–21.99)

Table 5 The order of oxidative stress in C. auratus livers caused by different treatments.

3. Results 3.1. The toxicity of TBBPA and TBBPA-D to D. magna

7 day

The result of the sensitivity test of D. magna (EC50 = 0.7 mg K2 Cr2 O7 /L) indicated that the D. magna strain used in this study met the requirement (EC50 = 0.5 ∼ 1.2 mg K2 Cr2 O7 /L) of the OECD guideline (OECD, 2004). The acute toxicity of TBBPA and TBBPA-D to D. magna is presented in Table 3. The EC50 values of TBBPA and TBBPA-D increased with increasing pH. The EC50 values of TBBPA-D were lower than those of TBBPA, except at pH 6.0, where the EC50 values were approximately equal for TBBPA and TBBPA-D. Additionally, linear relationships between the EC50 values of TBBPA and TBBPA-D and pH were identified. The plots of the linear relationships are shown in Supporting information (Fig. S1A and B). TBBPAEC50 = 8.80pH – 54.29(R2 = 0.96) TBBPA-DEC50 = 6.18pH – 40.18(R = 0.84).

(3)

3.2. The toxicity of TBBPA and TBBPA-D to L. hoffmeisteri The acute toxicity of TBBPA and TBBPA-D to L. hoffmeisteri is presented in Table 4. In general, the EC50 values increased with increasing pH, and the EC50 values of TBBPA were higher than those of TBBPA-D. However, the differences between the EC50 values of TBBPA and TBBPA-D for L. hoffmeisteri were smaller than those for D. magna. Similar to the results of the toxicity study in D. magna, TBBPA and TBBPA-D exhibited the most acute toxic effects at pH 6.0 in L. hoffmeisteri. Additionally, linear relationships between the EC50 values of TBBPA and TBBPA-D and pH were identified. The plots of the linear relationships are shown in Supporting information (Fig. S1C and D). TBBPAEC50 = 5.97pH – 34.13(R2 = 0.95)

(4)

TBBPA-DEC50 = 4.90pH – 28.48(R2 = 0.90).

(5)

Treatments

IBR values Treatments

IBR values

Untreated control (pH 7.5) pH (6.0) pH (7.5)–TBBPA (0.3) pH (6.0)–TBBPA (3) pH (9.0)–TBBPA (3) pH (9.0)–TBBPA (0.3) pH (6.0)–TBBPA (0.3) pH (9.0) pH (7.5)–TBBPA (3)

0.00 4.20 6.73 7.13 8.11 8.16 8.73 9.52 17.09

0.00 3.94 4.77 5.88 6.58 7.44 7.91 7.92 12.17

Untreated control (pH 7.5) pH (6.0) pH (7.5)–TBBPA (0.3) pH (9.0)–TBBPA (0.3) pH (6.0)–TBBPA (0.3) pH (6.0)–TBBPA (3) pH (9.0) pH (9.0)–TBBPA (3) pH (7.5)–TBBPA (3)

control. Additionally, with increasing TBBPA concentrations in the pH (9.0)-treated groups, there was a slight increasing trend in CAT activity after 7 day of exposure (Fig. 1B) and, in contrast, a noteworthy decreasing trend after 21 day of exposure (Fig. 2B). GSH level in C. auratus following co-exposure to TBBPA and pH is presented in Figs. 1 and 2D. After 7 day of exposure, GSH level was significantly decreased (p < 0.05) in all treatment groups compared with the untreated control. After 21 day of exposure, increasing trends in GSH level with TBBPA concentrations were observed for all the treated groups.

(2)

2

21 day

3.4. Integrated biomarker response The star plots for biomarkers were shown in Fig. 3. The calculated IBR values of all groups after 7 day and 21 day of exposures are presented in Table 5. According to this index, after 7 day or 21 day of exposures, the fish livers of the pH (7.5)–TBBPA (3) group experienced the most serious stress, and the least stressful condition was the untreated control. There was different order of oxidative stress after 7 day and 21 day of exposure. The IBR values after 7 day of exposure were general higher than its corresponding values after 21 day of exposure, except for the pH (6.0)–TBBPA (3) groups.

3.3. Oxidative stress in fish

3.5. Histopathological study of C. auratus gills and livers

Changes in the measured oxidative stress biomarkers, including changes in the activity of the antioxidant enzymes SOD, CAT, and GST and changes in the level of the non-enzyme antioxidant GSH are presented in Figs. 1 and 2. After 7 day of exposure (Fig. 1), the pH-treated and TBBPAtreated groups showed significant decreases (p < 0.05) in SOD, CAT, and GST activity compared to the untreated control. However, after 21 day of exposure (Fig. 2), there were no significant differences for the SOD activity among the untreated control, pH (6.0)–TBBPA (0.3) and pH (7.5)–TBBPA (0.3) groups. SOD activity of the pH (9.0) group significantly increased as compared to the untreated control. Nevertheless, significant decreases in CAT and GST activity were observed in all the treatments compared with the untreated

To more clearly understand the toxicity of TBBPA, gills and livers of fish exposed to high TBBPA concentrations (i.e., the pH (6.0), pH (6.0)–TBBPA (3), untreated control, pH (7.5)–TBBPA (3), pH (9.0), and pH (9.0)–TBBPA (3) groups) were examined histologically after 21 day of exposure. The results are shown in Figs. 4 and 5. The following types of tissue damages were observed in the gills of fish: epithelial hyperplasia (eh); lamellar aneurism (la); mesothelioma (me); curling of lamella (cl) and shortening of lamella (sl) (Fig. 4A–F). Lesions caused by the TBBPA were more serious with contrast to pH-related groups. For example, there were only morphological changes (“eh” in Fig. 4B and “cl” in Fig. 4E) or slight tumors (“la” in Fig. 4C) in pH-related groups. However, in doserelated groups, morphological changes accompanied by cancer

Table 4 The EC50 values of the toxicity of TBBPA and TBBPA dissolved by DMSO to L. hoffmeisteri with confidence limits (95% probability), n = 3. pH

6.0

7.0

7.5

8.0

9.0

TBBPA TBBPA + DMSO

2.92 (2.28–3.65) 3.20 (2.51–4.35)

3.59 (2.42–5.55) 6.49 (4.29–9.61)

15.26 (10.92–23.57) 6.62 (5.21–8.22)

12.16 (9.37–15.36) 11.75 (10.06–13.96)

21.22 (16.54–26.25) 16.54 (14.37–18.96)

150

Q. He et al. / Aquatic Toxicology 164 (2015) 145–154

Fig. 1. The effects of TBBPA on the activitiy of SOD (A), CAT (B), and GST (C) as well as GSH level (D) in fish livers under different pH conditions after 7 day of exposure. Enzyme activity is expressed as U/mg protein, while GSH level is expressed as ␮mol/g protein. Data are means ± SD (n = 4). Values that do not share the same superscript letter (a–g) were significantly different (p < 0.05). In the figure, the pH (6.0) group is defined as the group exposed to a solution with pH 6.0 and a concentration of TBBPA of 0.00 mg/L. For further explanation, the pH (6.0)–TBBPA (0.3) group was exposed to a solution with pH 6.0 and a concentration of TBBPA of 0.3 mg/L, while the pH of the exposure solution and the concentration of TBBPA in the pH (6.0)–TBBPA (3) group were 6.0 and 3.0 mg/L, respectively. Similar expressions are used for the pH (7.5) and pH (9.0)-treated groups. The pH (7.5) group is defined as the control.

Fig. 2. The effects of TBBPA on the activity of SOD (A), CAT (B), and GST (C) as well as GSH level (D) in fish livers under different pH conditions after 21 day of exposure. Data are means ± SD (n = 4). Values that do not share the same superscript letter (a–g) were significantly different (p < 0.05). The definitions of the group names are presented in Fig. 1.

7d

CAT

CAT

4.00

4.00

Untreated control (pH 7.5)

3.00

pH (7.5)-TBBPA (0.3)

2.00

2.00

pH (7.5)-TBBPA (3)

1.00

1.00

3.00

SOD

21 d

GSH

0.00

SOD

pH (6.0) GSH

0.00

pH (6.0)-TBBPA (0.3) pH (6.0)-TBBPA (3) pH (9.0) pH (9.0)-TBBPA (0.3)

GST

pH(9.0)-TBBPA (3) GST

Fig. 3. Integrated biomarker response (IBR) indices of all of the measured parameters in the livers of C. auratus following co-exposure to TBBPA and pH. The definitions of the group names are presented in Fig. 1.

Q. He et al. / Aquatic Toxicology 164 (2015) 145–154

151

Fig. 4. Histology of gills in C. auratus in six experimental groups after 21 day of exposure (n = 3). (A) control showing normal appearance of gill filaments (f) and lamellae (l); (B) pH (7.5)–TBBPA (3); (C) pH (6.0); (D) pH (6.0)–TBBPA (3); (E) pH (9.0); (F) pH (9.0)–TBBPA (3). Epithelial hyperplasia (eh); lamellar aneurisms (la); mesothelioma (me); curling of lamella (cl); shortening of lamella (sl). The definitions of the group names are presented in Fig. 1.

were occurred (“me” in Fig. 4D and F). So, we may infer that TBBPA may induce cancer in fish gills. In the livers of these fish, pyknotic nuclei (pn), hyalinization (hy) and deformation of nuclei (dn) were commonly observed (Fig. 5A–F). There was no significant difference in the occurrence of histological lesions between pH-related and dose-related groups.

4. Discussion TBBPA has two polar hydroxyl groups and three different forms, including the molecular form (TBBPA), which exists in acidic conditions and two dissociated forms (TBBPA− and TBBPA2− ), which exist in alkaline conditions. As an ionic compound, its solubility and forms can be significantly influenced by pH (Li et al., 2013; Mao et al., 2010). Based on the dissociation equations mentioned by Kuramochi et al. (2008), the form distribution of TBBPA under different pH conditions is shown in Fig. 6. Generally, different forms of TBBPA have different toxic potencies. As shown in Tables 2 and 3, TBBPA and TBBPA-D were more toxic in acidic conditions than in alkaline conditions. In other words, molecular TBBPA was more toxic than dissociated forms. Because in acidic conditions, molecular TBBPA was predominant (Fig. 6) and it can interact with phospholipid membranes to distribute throughout all regions of the phospholipid bilayer to influence biological processes involving cell membranes (Ogunbayo et al., 2007).

In this study, TBBPA-D exhibited a higher toxic potency to D. magna and L. hoffmeisteri than TBBPA, which may be the result of DMSO. However, it was found no apparent impact when DMSO was administered alone to D. magna and L. hoffmeisteri in this study. Similarly, the effects of DMSO on different organisms have been studied in other researches (Hallare et al., 2006; Chen et al., 2011; David et al., 2012). For example, in zebrafish embryos, Hallare et al. (2006) did not observe an obvious effect on survival but did observe an induction of stress proteins after exposure to 0.1% (v/v) DMSO. The phenomenon in our study could be explained by the potential synergistic effect between TBBPA and DMSO. Specifically, DMSO is a lipophilic molecule and is thought to be relatively highly permeable throughout the cell membrane (Glass et al., 2006). In this regard, DMSO could allow the TBBPA molecule to enter into the cell more easily, resulting in the higher observed toxicity for TBBPA-D than for TBBPA. Environmental contaminants may trigger oxidative stress in aquatic organisms when antioxidant defenses are overcome by pro-oxidant forces. Oxidative damages are caused by an imbalance between the production and elimination of ROS. A variety of molecules are present in aquatic organisms to protect them from oxidative damages, including water soluble reductants, such as glutathione; fat soluble vitamins, such as ␣-tocopherol; and enzymes, such as SOD, CAT, and GST (Digiulio et al., 1989; Kappus, 1986).

152

Q. He et al. / Aquatic Toxicology 164 (2015) 145–154

Fig. 5. Histology of livers in C. auratus in six experimental groups after 21 day of exposure (n = 3). (A) control showing normal hepatocytes; (B) pH (7.5)–TBBPA (3); (C) pH (6.0); (D) pH (6.0)–TBBPA (3); (E) pH (9.0); (F) pH (9.0)–TBBPA (3). Pyknotic nuclei (pn); hyalinization (hy); deformation of nuclei (dn). The definitions of the group names are presented in Fig. 1.

Percentage of three forms of TBBPA (%)

100

80

TBBPA TBBPA 2TBBPA

60

40

20

0 6

7

8

pH

9

10

11

Fig. 6. The percentages of three forms of TBBPA at different pH values.

SOD and CAT are considered as the first line of defense against ROS, and they act as ROS scavengers (Pandey et al., 2003). SOD catalyzes the dismutation of superoxide to H2 O2 , which is further detoxified by CAT (Hu et al., 2009). In the present study, after 7 day of exposure to TBBPA, SOD activity was significantly lower than in the untreated control. This remarkable reduction was presum-

ably due to the superfluous production of ROS, which exceeded the ROS removal capacity of SOD (Feng et al., 2013a). In this work, after 7 or 21 day of exposure to TBBPA, CAT activity was strongly inhibited in almost all of the treatment groups as compared to the untreated control. This suggests that the formation and accumulation of H2 O2 exceeded the antioxidant capacity of CAT, leading to the observed decrease in CAT activity (Xue et al., 2009). GST, a crucial phase II detoxification enzyme, is involved in the catalysis of the conjugation of GSH to electrophilic metabolites, making them more water soluble and easier to be excreted in the urine or bile (Feng et al., 2013b). TBBPA contains an electrophilic hydroxyl radical. It can directly combine with GST to form intermediate metabolites that can be routed to the bile, resulting in less accumulation of TBBPA in the livers (Yang et al., 2013). Through this combining function, GST protects livers from damages caused by TBBPA; thus, GST activity may reflect chronic TBBPA exposure to some degree. In this research, GST activity was substantially decreased after TBBPA exposure when compared to the untreated control. Depending on the TBBPA dose, exposure time and the susceptibility of the exposed species, enhancement (Napierska et al., 2009), diminishment (Feng et al., 2013a), and even enhancement followed by diminishment (Sun et al., 2006; Yang et al., 2013) of GST activity can occur in fish livers (van der Oost et al., 2003). The catalysis of GSH to electrophilic metabolites to livers may contribute to the GST decline

Q. He et al. / Aquatic Toxicology 164 (2015) 145–154

after TBBPA exposure (Yang et al., 2013). However, in this study, GST activity showed varying responses to increasing TBBPA concentrations; thus, more studies are needed to more fully understand this phenomenon. GSH is a major thiol containing molecule in cells; it is the predominant defense against ROS and acts by protecting cells from damages caused by electrophilic compounds (Oliveira et al., 2009). Changes in GSH level may indicate the detoxification ability of a test species. A higher GSH level indicates a greater ability of the cell to destroy free radicals (Cheung et al., 2001). Many previous studies claimed that GSH level was decreased after exposure to waterborne pollutants. Feng et al. (2003b) showed that GSH level was significantly decreased in a time-dependent manner in C. auratus livers after exposure to TBBPA. Zhang et al. (2008) also observed remarkable decrease in GSH level in livers of Gobiocypris rarus exposed to hexabromocyclododecane (HBCDD). However, in this study, the GSH levels in most of the TBBPA-treated groups were shown to increase in both a time- and dose-dependent manner. These increases in GSH levels after TBBPA exposure were in accordance with the results of Stephensen et al. (2002) and Mulcahy et al. (1997) and can most likely be explained by the up-regulation of enzymes that participate in GSH synthesis (Stephensen et al., 2002). Many important functions take place in gills, including gas exchange, ionic and osmatic regulation, and acid–base equilibrium (Song et al., 2013). TBBPA in aquatic environments can be taken up through the gills and distributed to other organs, including liver which is a major detoxification organ. In this study, histological examinations revealed some tissue damages, e.g., epithelial hyperplasia and lamellar aneurisms in gills; pyknotic nuclei and hyalinization in livers. Moreover, dose-related groups caused more serious lesions than pH-related groups. This result could provide more evidences for toxicological impacts of TBBPA on gills. The lamellar epithelium lifting and hyperplasia could serve as a defense function, because these histological changes could increase the distance across which waterborne irritants must diffuse to reach the bloodstream (Song et al., 2013). However, Kuiper et al. (2007a) did not observe any histological changes in the livers or gills of European flounders after TBBPA exposure. This may attribute to differences in the TBBPA doses and the sensitivities of the test organisms used in these two studies. Furthermore, the IBR results showed that fish exposed to tap water with a high TBBPA concentration (i.e., pH (7.5)–TBBPA (3)) suffered the most oxidative stress. The IBR values after 7 day of exposure were general higher than its corresponding after 21 day of exposure, except for the pH (6.0)–TBBPA (3) groups. This is because that TBBPA not only induces lipid peroxidation in liver tissues, but also activates the circulatory phagocytes, which further contribute to the increased burden of ROS in fish (Ahmad et al., 2000). So in this study, after 21 day of exposure, more circulatory phagocytes were activated to enhance the antioxidative ability of C. auratus leading to smaller IBR values after 21 day of exposure.

5. Conclusion TBBPA was investigated for its low-dose, acute, and chronic toxic effects on aquatic organisms in more realistic aquatic environmental conditions. It could be concluded that TBBPA was more toxic to D. magna and L. hoffmeisteri in acidic conditions, strong linear relationships were observed between the EC50 values of TBBPA and pH values. Additionally, there is synergistic effect between DMSO and TBBPA. The oxidative stress study showed that the activity of SOD, CAT, and GST and the GSH levels in all treated groups were altered by TBBPA exposure, which indicates that TBBPA enhanced the production of ROS and induced the oxidative stress in C. auratus

153

livers. Furthermore, the IBR values provided the order of oxidative stress in C. auratus livers after TBBPA exposure. Histological study of gills and livers supplied more evidences of the toxicological effects caused by TBBPA. In summary, this work provides some basic information for TBBPA risk assessment and supplements for previously published data on the toxic mechanisms of TBBPA. Further investigations will be focused on relationships between histological changes and toxicological mechanisms of TBBPA. Acknowledgements This research was financially supported by the National Natural Science Foundation of China (No. 41071319, 21377051), the Major Science and Technology Program for Water Pollution Control and Treatment of China (No. 2012ZX07506-001) and the Scientific Research Foundation of Graduate School of Nanjing University (2013CL08). Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.aquatox. 2015.05.005. References Ahmad, I., Hamid, T., Fatima, M., Chand, H.S., Jain, S.K., Athar, M., Raisuddin, S., 2000. Induction of hepatic antioxidants in freshwater catfish (Channa punctatus Bloch) is a biomarker of paper mill effuent exposure. Biochim. Biophys. Acta 1523, 37–48, http://dx.doi.org/10.1016/S0304-4165(00)98-2. Birnbaum, L.S., Staskal, D.F., 2003. Brominated flame retardants: cause for concern? Environ. Health Perspect. 112, 9–17, http://dx.doi.org/10.1289/ehp.6559. Bradford, M.M., 1976. A rapid and sensitive method for the quantitation of microgram quantities of protein utilizing the principle of protein-dye binding. Anal. Biochem. 72, 248–254, http://dx.doi.org/10.1016/0003-2697(76) 90527-3. Carlsson, G., Norrgren, L., 2014. Comparison of embryo toxicity using two classes of aquatic vertebrates. Environ. Toxicol. Pharmacol. 37, 24–27, http://dx.doi.org/10.1016/j.etap.2013.10.015. Chen, T.H., Wang, Y.H., Wu, Y.H., 2011. Developmental exposures to ethanol or dimethylsulfoxide at low concentrations alter locomotor activity in larval zebrafish: implications for behavioral toxicity bioassays. Aquat. Toxicol. 102, 162–166, http://dx.doi.org/10.1016/j.aquatox.2011.01.010. Cheung, C.C.C., Zheng, G.J., Li, A.M.Y., Richardson, B.J., Lam, S., 2001. Relationships between tissue concentrations of polycyclic aromatic hydrocarbons and antioxidative responses of marine mussels, Perna viridis. Aquat. Toxicol. 52, 189–203, http://dx.doi.org/10.1016/S0166-445X(00). Colnot, T., Kacew, S., Dekant, W., 2014. Mammalian toxicology and human exposures to the flame retardant 2,2 ,6,6 -tetrabromo-4,4 -isopropylidenediphenol (TBBPA): implications for risk assessment. Arch. Toxicol. 88, 553–573, http://dx.doi.org/10.1007/s00204-013-1180-8. Damiens, G., Barelli, M.G., Loquès, F., Roméo, M., Salbert, V., 2007. Integrated biomarker response index as a useful tool for environmental assessment evaluated using transplanted mussels. Chemosphere 66, 574–583, http://dx.doi.org/10.1016/j.chemosphere.2006.05.032. David, R.M., Jones, H.S., Panter, G.H., Winter, M.J., Hutchinson, T.H., Chipman, J.K., 2012. Interference with xenobiotic metabolic activity by the commonly used vehicle solvents dimethylsulfoxide and methanol in zebrafish (Danio rerio) larvae but not Daphnia magna. Chemosphere 88, 912–917, http://dx.doi.org/10.1016/j.chemosphere.2012.03.018. ECB, 2006. European union risk assessment report—2,2 ,6,6 -tetrabromo-4,4 -isopropylidenediphenol (tetrabromobisphenol-A or TBBP-A) (CAS: 79-94-7) Part II—human health, vol 63, EUR 22,161 EN. Institute for Health and Consumer Protection, European Chemicals Bureau, European Commission Joint Research Centre, 4th Priority List, Luxembourg: Office for Official Publications of the European Communities. ECB, 2008. European Union risk assessment report—2,2 ,6,6 -tetrabromo-4,4 -isopropylidenediphenol (tetrabromobisphenol-A or TBBP-A) (CAS: 79-94-7) Part I—environment (final draft). Decherf, S., Demeneix, B.A., 2011. The obesogen hypothesis: a shift of focus from the periphery to the hypothalamus. J. Toxicol. Environ. Health B Crit. Rev. 14, 423–448, http://dx.doi.org/10.1080/10937404.2011.578561. de Schamphelaere, K.A.C., Heijerick, D.G., Janssen, C.R., 2004. Comparison of the effect of different pH buffering techniques on the toxicity of copper and zinc to Daphnia Magna and Pseudokirchneriella Subcapitata. Ecotoxicology 13, 697–705, http://dx.doi.org/10.1007/s10646-003-4429-9.

154

Q. He et al. / Aquatic Toxicology 164 (2015) 145–154

de Wit, C.A., 2002. An overview of brominated flame retardants in the environment. Chemosphere 46, 583–624, http://dx.doi.org/10.1016/S0045-6535(01)225-9. de Wit, C.A., Herzke, D., Vorkamp, K., 2010. Brominated flame retardants in the Arctic environment-trends and new candidates. Sci. Total Environ. 408, 2885–2918, http://dx.doi.org/10.1016/j.scitotenv.2009.08.037. Digiulio, R.T., Washburn, P.C., Wenning, R.J., Winston, G.W., Jewell, C.S., 1989. Biochemical responses in aquatic animals: a review of determinants of oxidative stress. Environ. Toxicol. Chem. 8, 1103–1123, http://dx.doi.org/10.1002/etc.5620081203. Feng, M.B., Li, Y., Qu, R.J., Wang, L.S., Wang, Z.Y., 2013a. Oxidative stress biomarkers in freshwater fish Carassius auratus exposed to decabromodiphenyl ether and ethane, or their mixture. Ecotoxicology 22, 1101–1110, http://dx.doi.org/10.1007/s10646-013-1097-2. Feng, M.B., Qu, R.J., Wang, C., Wang, L.S., Wang, Z.Y., 2013b. Comparative antioxidant status in freshwater fish Carassius auratus exposed to six current-use brominated flame retardants: a combined experimental and theoretical study. Aquat. Toxicol. 140, 314–323, http://dx.doi.org/10.1016/j.aquatox.2013.07.001. Góth, L., 1991. A simple method for determination of serum catalase activity and revision of reference range. Clin. Chim. Acta 196, 143–151, http://dx.doi.org/10.1016/0009-8981(91)90067-M. Glass, C.A., Perrin, R.M., Pocock, T.M., Bates, D.O., 2006. Transient osmotic absorption of fluid in microvessels exposed to low concentrations of dimethyl sulfoxide. Microcirculation 13, 29–40, http://dx.doi.org/10.1080/10739680500383464. Hallare, A., Nagel, K., Kohler, H.R., Triebskorn, R., 2006. Comparative embryotoxicity and proteotoxicity of three carrier solvents to zebrafish (Danio rerio) embryos. Ecotox. Environ. Safe. 63, 378–388, http://dx.doi.org/10.1016/j.ecoenv.2005.07.006. Hamers, T., Kamstra, J.H., Sonneveld, E., Murk, A.J., Kester, M.H., Andersson, P.L., Legler, J., Brouwer, A., 2006. In vitro profiling of the endocrine-disrupting potency of brominated flame retardants. Toxicol. Sci. 92, 157–173, http://dx.doi.org/10.1093/toxsci/kfj187. Hu, F.X., Pan, L.Q., Xiu, M., Liu, D., 2015. Dietary accumulation of tetrabromobisphenol A and its effects on the scallop Chlamys farreri. Comp. Biochem. Physiol. C 167, 7–14, http://dx.doi.org/10.1016/j.cbpc.2014.08.002. Hu, J., Liang, Y., Chen, M.J., Wang, X.R., 2009. Assessing the Toxicity of TBBPA and HBCD by zebrafish embryo toxicity assay and biomarker analysis. Environ. Toxicol. 24, 334–342, http://dx.doi.org/10.1002/tox.20436. Jollow, D., Mitchell, J., Zampaglione, N.A., Gillette, J., 1974. Bromobenzene-induced liver necrosis. Protective role of glutathione and evidence for 3,4-bromobenzene oxide as the hepatotoxic metabolite. Pharmacology 11, 151–169, http://dx.doi.org/10.1159/000136485. Kappus, H., 1986. Overview of enzyme-systems involved in bio-reduction of drugs and in redox cycling. Biochem. Pharmacol. 35, 1–6, http://dx.doi.org/10.1016/0006-2952(86)90544-7. Kim, W.K., Lee, S.K., Jung, J., 2010. Integrated assessment of biomarker responses in common carp (Cyprinus carpio) exposed to perfluorinated organic compounds. J. Hazard. Mater. 180, 395–400, http://dx.doi.org/10.1016/j.jhazmat.2010.04.044. Kuester, R.K., Sólyom, A.M., Rodriguez, V.P., Sipes, I.G., 2007. the effects of dose, route, and repeated dosing on the dispositionand kinetics of tetrabromobisphenol a in male F-344 rats. Toxicol. Sci. 96, 237–245, http://dx.doi.org/10.1093/toxsci/kfm006. Kuiper, R.V., Canton, R.F., Leonards, P.E., Jenssen, B.M., Dubbeldam, M., Wester, P.W., van den Berg, M., Vos, J.G., Vethaak, A.D., 2007a. Long-term exposure of European flounder (Platichthys flesus) to the flame-retardants tetrabromobisphenol A (TBBPA) and hexabromocyclododecane (HBCD). Ecotoxicol. Environ. Safe. 67, 349–360, http://dx.doi.org/10.1016/j.ecoenv.2006.12.001. Kuiper, R.V., van den Brandhof, E.J., Leonards, P.E., van der Ven, L.T., Wester, P.W., Vos, J.G., 2007b. Toxicity of tetrabromobisphenol A (TBBPA) in zebrafish (Danio rerio) in a partial life-cycle test. Arch. Toxicol. 81, 1–9, http://dx.doi.org/10.1007/s00204-006-0117-x. Kuramochi, H., Kawamoto, K., Miyazaki, K., Nagahama, K., Maeda, K., Li, X.W., Shibata, E., Nakamura, T., Sakai, S.I., 2008. Determination of physicochemical properties of tetrabromobisphenol A. Environ. Toxicol. Chem. 27, 2413–2418, http://dx.doi.org/10.1897/07-472.1. Li, J., Bai, J., Zhou, B., Hu, X., 2013. Solubility of 2,2 ,6,6 -tetrabromo-4,4 -isopropylidene phenol in aqueous pollutant solutions. J. Chem. Eng. Data 58, 3150–3154, http://dx.doi.org/10.1021/je400602s. Mao, L., Yu, Y.J., Sun, Z.H., Yu, H.X., 2010. Water solubility and octanol/water partition behaviour of tetrabromobisphenol A. Environ. Chem. 4, 006. McCord, J.M., Fridovich, I., 1969. Superoxide dismutase an enzymic function for erythrocuprein (hemocuprein). J. Biol. Chem. 244, 6049–6055. Morris, S., Allchin, C.R., Zegers, B.N., Haftka, J.J.H., Boon, J.P., Belpaire, C., Leonards, P.E.G., van Leeuwen, S.P.J., De Boer, J., 2004. Distributon and fate of HBCD and TBBPA brominated flame retardants in north sea estuaries and aquatic food webs. Environ. Sci. Technol. 38, 5497–5504, http://dx.doi.org/10.1021/es049640i. Mulcahy, R.T., Wartman, M.A., Bailey, H.H., Gipp, J.J., 1997. Constitutive and ␤-naphthoflavone-induced expression of the human ␥-glutamylcysteine synthetase heavy subunit gene is regulated by a distal antioxidant response

element/TRE sequence. J. Biol. Chem. 272, 7445–7454, http://dx.doi.org/10.1074/jbc.272.11.7445. Napierska, D., Barsiene, J., Mulkiewicz, E., Podolska, M., Rybakovas, A., 2009. Biomarker responses in flounder Platichthys flesus from the Polish coastal area of the Baltic Sea and applications in biomonitoring. Ecotoxicology 18, 846–859, http://dx.doi.org/10.1007/s10646-009-0328-z. OECD, 2004. OECD guideline for testing of chemicals—Daphnia sp., acute immobilisation test 202. In: Effects on Biotic Systems. OECD, Section 2. Ogunbayo, O.A., Jensen, K.T., Michelangeli, F., 1768. The interaction of the brominated flame retardant: tetrabromobisphenol A with phospholipid membranes. Biochim. Biophys. Acta 155, 1559–1566, http://dx.doi.org/10.1016/j.bbamem.2007.03.013. Oliveira, M., Maria, V.L., Ahmad, I., Serafim, A., Bebianno, M.J., Pacheco, M., Santos, M.A., 2009. Contamination assessment of a coastal lagoon (Ria de Aveiro, Portugal) using defence and damages biochemical indicators in gill of Liza aurata – an integrated biomarker approach. Environ. Pollut. 157, 959–967, http://dx.doi.org/10.1016/j.envpol.2008.10.019. Osako, M., Kim, Y.J., Sakai, S.I., 2004. Leaching of brominated flame retardants in leachate from landfills in Japan. Chemosphere 57, 1571–1579, http://dx.doi.org/10.1016/j.chemosphere.2004.08.076. Pandey, S., Parvez, S., Sayeed, I., Haque, R., Bin-Hafeez, B., Raisuddin, S., 2003. Biomarkers of oxidative stress: a comparative study of river Yamuna fish Wallago attu (Bl. & Schn.). Sci. Total Environ. 309, 105–115, http://dx.doi.org/10.1016/S0048-9697(03)6-8. Reistad, T., Mariussen, E., Ring, A., Fonnum, F., 2006. In vitro toxicity of tetrabromobisphenol A on cerebellar granule cells: cell death, free radical formation, calcium influx and extracellular glutamate. Toxicol. Sci. 96, 268–278, http://dx.doi.org/10.1093/toxsci/kfl198. Ronisz, D., Finne, E.F., Karlsson, H., Förlin, L., 2004. Effects of the brominated flame retardants hexabromocyclododecane (HBCDD), and tetrabromobisphenol A (TBBPA), on hepatic enzymes and other biomarkers in juvenile rainbow trout and feral eelpout. Aquat. Toxicol. 69, 229–245, http://dx.doi.org/10.1016/j.aquatox.2004.05.007. Schauer, U.M.D., Völkel, W., Dekant, W., 2006. Toxicokinetics of tetrabromobisphenol a in humans and rats after oral administration. Toxicol. Sci. 91, 49–58, http://dx.doi.org/10.1093/toxsci/kfj132. Sellstrom, U., Jansson, B., 1995. Analysis of tetrabromobisphenol A in a product and environmental-samples. Chemosphere 31, 3085–3092, http://dx.doi.org/10.1016/0045-6535(95)167-7. Song, Y.F., Luo, Z., Huang, C., Liu, X., Pan, Y.X., Chen, Q.L., 2013. Effects of calcium and copper exposure on lipogenic metabolism, metal element compositions and histology in Synechogobius hasta. Fish Physiol. Biochem. 39, 1641–1656, http://dx.doi.org/10.1007/s10695-013-9816-4. Stephensen, E., Sturve, J., Forlin, L., 2002. Effects of redox cycling compounds on glutathione content and activity of glutathione-related enzymes in rainbow trout liver. Comp. Biochem. Physiol. C Toxicol. Pharmacol. 133, 435–442, http://dx.doi.org/10.1016/S1532-0456(02)129-1. Sun, Y.Y., Yu, H.X., Zhang, J.F., Yin, Y., Shen, H., Liu, H.L., Wang, X.R., 2006. Bioaccumulation and antioxidant responses in goldfish Carassius auratus under HC orange No. 1 exposure. Ecotoxicol. Environ. Safe. 63, 430–437, http://dx.doi.org/10.1016/j.ecoenv.2005.02.001. Sun, Z.H., Yu, Y.J., Mao, L., Zheng, F., Yu, H.X., 2008. Sorption behaviour of tetrabromobisphenol A in two soils with different characteristics. J. Hazard. Mater. 160, 450–461, http://dx.doi.org/10.1016/j.jhazmat.2008.03.019. van Boxtel, A.L., Kamstra, J.H., Cenijn, P.H., Pieterse, B., Wagner, J.M., Antink, M., Krab, K., van der Burg, B., Marsh, G., Brouwer, A., Legler, J., 2008. Microarray analysis reveals a mechanism of phenolic polybrominated diphenylether toxicity in zebrafish. Environ. Sci. Technol. 42, 1773–1779, http://dx.doi.org/10.1021/es0720863. van der Oost, R., Beyer, J., Vermeulen, N.P., 2003. Fish bioaccumulation and biomarkers in environmental risk assessment: a review. Environ. Toxicol. Pharmacol. 13, 57–149, http://dx.doi.org/10.1016/S1382-6689(02)126-6. Watanabe, I., Kashimoto, T., Tatsukawa, R., 1983. Identification of the flame-retardant tetrabromobisphenol A in the river sediment and the mussel collected in osaka. Bull. Environ. Contam. Toxicol. 31, 48–52, http://dx.doi.org/10.1007/B1619835. WHO/ICPS, 1995. Environmental Health Criteria 172, Tetrabromobisphenol A and Derivatives. World Health Organization, Geneva, Switzerland http://www.inchem.org/documents/ehc/ehc/ehc172.htm Xue, Y.G., Gu, X.Y., Wang, X.R., Sun, C., Xu, X.H., Sun, J., Zhang, B.G., 2009. The hydroxyl radical generation and oxidative stress for the earthworm Eisenia fetida exposed to tetrabromobisphenol A. Ecotoxicology 18, 693–699, http://dx.doi.org/10.1007/s10646-009-0333-2. Yang, S.W., Yan, Z.G., Xu, F.F., Wang, S.R., Wu, F.C., 2012. Development of freshwater aquatic life criteria for tetrabromobisphenol A in China. Environ. Pollut. 169, 59–63, http://dx.doi.org/10.1016/j.envpol.2012.05.023. Yang, S.W., Xu, F.F., Zheng, B.H., Wu, F.C., Wang, S.R., 2013. Multibiomarker responses upon exposure to tetrabromobisphenol A in the freshwater fish Carassius auratus. Aquat. Toxicol. 142, 248–256, http://dx.doi.org/10.1016/j.aquatox.2013.08.013. Zhang, X., Yang, F.X., Zhang, X.L., Xu, Y., Liao, T., Song, S.B., Wang, J.W., 2008. Induction of hepatic enzymes and oxidative stress in Chinese rare minnow (Gobiocypris rarus) exposed to waterborne hexabromocyclododecane (HBCDD). Aquat. Toxicol. 86, 4–11, http://dx.doi.org/10.1016/j.aquatox.2007.07.002.