Ecotoxicology and Environmental Safety 119 (2015) 116–122
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Hepatic oxidative stress biomarker responses in freshwater fish Carassius auratus exposed to four benzophenone UV filters Hui Liu a,b, Ping Sun b, Hongxia Liu b, Shaogui Yang a, Liansheng Wang a, Zunyao Wang a,n a b
State Key Laboratory of Pollution Control and Resources Reuse, School of Environment, Nanjing University, Jiangsu Nanjing 210023, PR China College of Biological and Chemical Engineering, Jiaxing University, Zhejiang Jiaxing 314001, PR China
art ic l e i nf o
a b s t r a c t s
Article history: Received 15 October 2014 Received in revised form 10 May 2015 Accepted 11 May 2015
Benzophenone (BP) type UV filters are widely used in many personal care products to protect human from UV irradiation. Despite the estrogenic potencies to fish and the environmental occurrences of BP derivatives in aquatic systems, little information is available regarding their effects on the antioxidant defense system in fish. In this work, the oxidative stress induced 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. Moreover, liver tissues were also studied histologically. The changes in the activities of antioxidant enzymes and glutathione levels suggested that BPs generates oxidative stress in fish. The IBR index revealed that the hepatic oxidative toxicity followed the order BP-1 4BP-2 4BP4 4BP-3. The histopathological analysis revealed lesions caused by BPs. This investigation provides essential information for assessing the potential ecological risk of BP-type UV filters. & 2015 Elsevier Inc. All rights reserved.
Keywords: UV filters Benzophenones Oxidative stress Integrated biomarker response Histological analysis
1. Introduction Benzophenones (BPs) are one of the primary components in the UV filters family and are, extensively used as sunscreen agents in personal care products (PCPs) for the protection of skin and hair from UV irradiation (Chisvert et al., 2012). In the European Union, 2-hydroxyl-4-methoxyl benzophenone (BP-3) and 2-hydroxyl-4methoxyl benzophenone-5-sulfonic acid (BP-4) are approved to be used as UV filters in sunscreens at maximum individual concentrations of 10% and 5%, respectively, and in Japan, 2,4-dihydroxybenzophenone (BP-1), 2,2′,4,4′-tetrahydroxybenzophenone (BP-2) and 2,2′-dihydroxy-4,4′-dimethoxybenzophenone (BP-6) are also allowed in sunscreens. Currently, BPs have been classified as chemicals suspected of having endocrine disrupting effects due to increasing evidence that BPs are able to interfere with the endocrine system (Kerdivel et al., 2013). BPs have been detected in environmental matrices such as water (Almeida et al., 2013; Wu et al., 2013; Jurado et al., 2014), soil, sediment (Zhang et al., 2011) and indoor dust (Wang et al., 2013a). The amounts found in water samples are in the ng L 1 range, which are not far below the doses that cause toxic effects in animals (Tarazona et al., 2010). Fish are important organisms to monitor the occurrence of persistent lipophilic contaminants (Fent et al., 2010). The presence of BPs UV filters was reported in fish n
Corresponding author. Fax: þ86 25 89680358. E-mail address: [email protected] (Z. Wang).
http://dx.doi.org/10.1016/j.ecoenv.2015.05.017 0147-6513/& 2015 Elsevier Inc. All rights reserved.
from Switzerland rivers, Swiss lakes and Texas streams (Balmer et al., 2005; Mottaleb et al., 2009; Fent et al., 2010), even in marketed fish (Tsai et al., 2014). Some BP UV-filters have hormonal activity in fish. It is reported that BP-1, BP-2, BP-3 and BP-4 elicit estrogenic and anti-androgenic activities in fish (Kunz et al., 2006) in vivo and in vitro tests, and it is noteworthy that BP-1, BP-2 and 4-hydroxybenzophenone (4BP) possess estrogenic activities higher than that of BP-3 (Morohoshi et al., 2005). BP-2 could negatively affect egg production and gonadal development in fish (Weisbrod et al., 2007). In addition, induction of vitellogenin and impairment of reproduction in fish (Japanese medaka and rainbow trout), and alterations of gene expression in both adult zebrafish and eleuthero-embryos were reported with an aqueous exposure to BP-3 (Coronado et al., 2008; Bluthgen et al., 2012). Alterations of gene expression involved in hormonal pathways and steroidogenesis were also reported in zebrafish exposed to BP-4 (Zucchi et al., 2011). BPs have even been detected in human samples, such as breast milk, urine (Asimakopoulos et al., 2014a, 2014b) and blood (Zhang et al., 2013; Vela-Soria et al., 2014). Recently, through concentration analysis of BP-3 as well as four of its metabolic derivatives, BP1, BP-2, 2,2′-dihydroxyl-4-methoxl benzophenone (BP-8), and 4BP in urine of children and adults, BP-3 was found in nearly all urine samples from the U.S. and China (Wang and Kannan, 2013b). High urinary concentrations of BP-1 are associated with estrogen-dependent diseases, such as endometriosis in women (Kunisue et al., 2012). These findings demonstrate the apparent likelihood that BPs contained in PCPs might be emitted into the aquatic
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environment and persistent with a great potential for bioaccumulation through food chains (Gago-Ferrero et al., 2013). Thus, the occurrence (Liao and Kannan, 2014; Tsui et al., 2014), the degradation (Beel et al., 2013; Duirk et al., 2013; Yang and Ying, 2013; Ji et al., 2013; Laurentiis et al., 2013; Xiao et al., 2013) and, particularly the potential toxicities to humans and the ecosytems of these BP UV-filters have attracted widespread attention (Hofkamp et al., 2008; Downs et al., 2014; Paredes et al., 2014; Liu et al., 2015). Biomarkers such as enzyme activity, protein level and DNA can be used to measure the interaction between biological systems and physical, chemicals or biological environmental agents. And the evaluation of oxidative stress biomarkers is critical to the investigation of oxidative stress in organisms. Grabicova et al. found, after 42 days of UV filter 2-phenylbenzimidazole-5-sulfonicacid (PBSA) exposure in rainbow trout liver, the catalase (CAT) and superoxide dismutase (SOD) activities were not significantly different between the control group and the group that was exposed to the highest concentration of PBSA (1000 mg L 1), and no pathological changes were obvious in the liver or gonads (Grabicova et al., 2013). Gao et al. observed increased CAT activity in the Tetrahymena thermophile exposed to 1.0 μg L 1 BP-3, while no significant changes of the SOD activity (Gao et al., 2013). Therefore, we attempted to test the hypothesis whether these environmentally important benzophenones induce oxidative stress in fish. In the present study, four oxidation-related biomarkers, the activities of SOD, CAT and glutathione S-transferase (GST), and the level of glutathione (GSH) were measured to estimate the single toxicities of four common BP-type UV filters on antioxidant status in fish Carassius auratus (C. auratus). Next, integrated biomarker response (IBR) index was applied to improve their discriminatory power, and was calculated to estimate the integral effects of BPsinduced oxidative stress. Finally, liver tissues were examined histologically.
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Table 1 Water quality parameters of the water used for acclimation and subsequent experiments. Water quality parameters pH Conductivity Total hardness Alkalinity Na þ
7.25 7 0.25 350.67 12.5 μs/cm 146.5 7 9.3 mg CaCO3/L 40.5 75.0 mg CaCO3/L 12.2 70.2 mg/L
Kþ Mg2 þ Ca2 þ Cl DO (dissolved oxygen)
2.34 7 0.05 mg/L 7.75 7 0.02 mg/L 42.05 7 0.82 mg/L 28.17 1.2 mg/L 6.5 7 0.5 mg O2/L
water. The solubility of BP-3 is the most smallest than others, and the saturated solubility of BP-3 is 5.1 mg L 1 (Gago-Ferrero et al., 2013). Thus the lower concentration 0.5 mg L 1 were chose as close to worst-case environmental concentrations, the higher one of 5 mg L 1 as a pharmacological concentration. BPs were analyzed on an Agilent 1200 High Performance Liquid Chromatograph (HPLC) equipped with a Diode Array Detector. A Zorbax 300 SBC18 column (4.6 150 mm, 5 μm) was used for the separation at 30 °C. The mobile phase was 0.3% formic acid in water (A) and methanol (B) with an isocratic elution of 20:80 (v/v). The injection volume was 100 mL, the flow rate was set at 1 mL min 1, and the detection wavelength was 290 nm. The actual exposure concentrations of BP-1, BP-2, BP-3 and BP-4 in low dose groups are 0.48 70.03, 0.487 0.05, 0.48 70.03 and 0.497 0.03 mg L 1, respectively, and in high dose groups 4.707 0.47, 4.78 70.40, 4.767 0.32 and 4.90 70.40 mg L 1, respectively. Every two days, a half of water in the tanks was exchanged with fresh water containing the same concentrations of BPs to maintain the constant concentrations during the experiments. The final DMSO concentration was 0.005% in all tanks except the aqueous control. The protocols for fish maintaining, experimentation and sacrifice were approved by the Ethics Committee of Nanjing University (Qu et al., 2014).
2. Materials and methods
2.4. Sample preparation and biochemical assessments
2.1. Chemicals and reagents
4 fish were randomly sampled from each group on days 7, 14 and 28, and then killed by a blow to the head and dissected for liver in each treatment. Subsequently, the liver of each fish was placed on an ice plate and rinsed with cold physiological saline (0.9% NaCl), next weighed, and homogenized with cold physiological saline using an Ultra Turrax homogenizer (IKA, Germany). The homogenates were centrifuged by Eppendorf 5417R centrifuge (Eppendorf, Germany) at 4000 rpm at 4 °C for 15 min, to get the supernatants for further biochemical analysis. Four biomarkers, including the activities of SOD, CAT and GST, and GSH level, were analyzed. The SOD activity was measured based on the inhibition of cytochrome c caused by the superoxide radical (McCord and Fridovich, 1969). The CAT activity was evaluated by monitoring residual H2O2 absorbance at 405 nm (Góth, 1991). The GST activity was measured using GST assay kit by a spectrophotometer with 1-chloro-2,4-dinitrobenzene (CDNB) and glutathione-reduced (GSH) as substrates at 412 nm (Han et al., 2013). The GSH level was evaluated following the procedure of Jollow et al. (1974) with 5,5′dithiobis-2-nitrobenzoic acid (DTNB) as the reagent. DTNB was reduced by free sulfhydryl groups of GSH to form yellow-colored 5-thio-nitrobenzoic acid. The activities of antioxidant enzymes were normalized by total protein and represented as a relative percentage of the control. Total proteins were determined using the Bradford method (Bradford, 1976).
BP-1 (CAS no:131-56-6, 99% purity), BP-2 (CAS no:131-55-5, 99% purity), BP-3 (CAS no:131-57-7, 99% purity) and BP-4 (CAS no:4065-45-6, 98% purity) were purchased from J&K Co. (Shanghai, China). The stock solutions were prepared in dimethylsulfoxide (DMSO). All other reagents were of analytical purity. The commercial assay kits for the analysis of the oxidative stress biomarkers were purchased from Nanjing Jiancheng Bioengineering Institute (Nanjing, China). 2.2. 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. 2.3. Animals and experimental exposure C. auratus (weight: 28.84 73.35 g) were obtained commercially from a local supplier (Nanjing, China). Before the experiment, the healthy fish were acclimatized for at least 7 d in tanks containing dechlorinated and aerated tap water at 2472 °C. The fish were fed with the commercial fish food twice a day but fasted 24 h prior to biochemical analysis. During the acclimated period, the total mortality was less than 1%. After acclimatization, fifteen acclimated fish were transferred to each aquarium containing 30 L dechlorinated and aerated
2.5. Integrated biomarker response (IBR) A battery of biomarkers is often used to evaluate the effects of
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exposure to chemical contaminants and detect responses to environmental stress. Unfortunately, field application of biomarkers is subject to various constraints that can limit data acquisition and prevent the use of multivariate methods during statistical analysis. Therefore, a simple method is needed to summarize biomarker responses and simplify their interpretation in biomonitoring programs. Integrated biomarker response is a method to summarize the biomarker responses into one general “stress index”. It was applied to assess the potential toxicity of BPs to fish. The procedures for the IBR calculation are described here briefly: (1) data were calculated for mean and SD (standard deviation). (2) Data were standardized according to the formula Y ¼(X m)/S, where Y is the standardized data, X is the data of each biomarker response, m is the mean data of the biomarker, and S is the standard deviation of the biomarker. (3) Z was calculated as Z¼Y in the case of activation or Z¼ Y in the case of inhibition. The minimum value (Min) was obtained. (4) S was calculated as S ¼Zþ ❘Min❘, where S Z0 and ❘Min❘ is the absolute value. (5) Calculation of star plot areas by multiplying the obtained value of each biomarker (Si) with the next biomarker (Si þ 1) and dividing each calculation by 2. (6) The corresponding IBR value was obtained as
80 *
70
Histopathological examination of fish liver was performed using light microscopy. The liver was sampled, immobilized in 10% formaldehyde solution, wrapped in wax, and cut into slices of 4– 5 μm thickness using an ultra-thin semiautomatic microtome (ERM-3000, China). The slices were later stained with hematoxylin and counter stained with eosin. Next, the liver slices were examined under a light microscope (ZEISS Axio Imager A1, Germany) at a magnification of 400 and evaluated for changes in morphology of BPs exposed tissues compared to the aqueous control. 2.7. Statistical analysis Statistical analyses were conducted using the SPSS statistical package (ver. 16.0, SPSS Company, Chicago, USA). Whether the data distributions were normal was determined by the Shapiro–Wilk test and the homogeneity of variances was determined by the
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0 Aq ue ou sC on So tro lv l en tC on tro l BP -1 (0 .5 ) BP -1 (5 ) BP -2 (0 .5 ) BP -2 (5 ) BP -3 (0 .5 ) BP -3 (5 ) BP -4 (0 .5 ) BP -4 (5 )
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SOD activity (U/mg Protein)
2.6. Histopathology
7d 14d 28d
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IBR¼ {[(S1 S2)/2] þ[(S2 S3)/2] þ ⋯ þ[(Si Si þ 1)/2] }, when i¼ n, Si þ 1 ¼S1 (Beliaeff and Burgeot, 2002).
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10 0 Aq ue ou sC on So tro lv l en tC on tro l BP -1 (0 .5 ) BP -1 (5 ) BP -2 (0 .5 ) BP -2 (5 ) BP -3 (0 .5 ) BP -3 (5 ) BP -4 (0 .5 ) BP -4 (5 )
Aq ue ou sC on So tro lv l en tC on tro l BP -1 (0 .5 ) BP -1 (5 ) BP -2 (0 .5 ) BP -2 (5 ) BP -3 (0 .5 ) BP -3 (5 ) BP -4 (0 .5 ) BP -4 (5 )
0
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Fig. 1. The effects of BPs on the antioxidant enzymes activities (SOD, CAT and GST) and GSH level in liver of C. auratus. Data are means 7 SD, n ¼4 for each data point. *Different from control (p o 0.05). **Different from control (p o 0.01).
H. Liu et al. / Ecotoxicology and Environmental Safety 119 (2015) 116–122
Levene test. One-way analysis of variance (ANOVA) with the Dunnett’s test was used to analyze differences between the control and the experimental groups. The significant difference contains two grades, significant (p o0.05) and highly significant (p o0.01). Results are expressed as means 7 SD.
3. Results 3.1. Oxidative stress in fish No mortality was observed during the acclimatization and the exposure period to BPs, and there was no significant difference (p 40.05) among the blank and the DMSO control for all biomarkers during the exposure. Changes in the measured oxidative stress biomarkers, including changes in the activities of the antioxidant enzymes activities (SOD, CAT and GST) and changes in the level of the non-enzyme antioxidantare (GSH) are shown in Fig. 1A–D. After 7 d of exposure, there was a significantly decreasing trend in the SOD activity in all BPs-treated groups (p o0.01). After 14 or 28 d of exposure, the SOD activity increased gradually and returned to the blank level, and no significant changes were observed in SOD activity in the lower dose treated groups, except in the high concentration treated groups (Fig. 1A). Similarly, after 7 d of exposure, there was a noteworthy decreasing in CAT activity (p o0.05 or p o0.01) in all dose groups (Fig. 1B). With an increasing expose time, the CAT activity also returned to the blank level gradually, except in all BP-2 treated groups. After 14 or 28 d of exposure, there was a noteworthy decreasing trend (p o0.05 or p o0.01) in the BP-2 group (0.5), and after 28 d of exposure in the BP-2 group (5). Moreover, the CAT activity was induced significantly after 14 d of exposure to BP-4 (0.5).
Fig. 1C shows the activity of GST during the exposure. GST activity increased by BP-1 and BP-3 significantly (po 0.05 or po 0.01) at first and peaked after 7 d of exposure, then remained after 14 d of exposure, but returned to the blank level after 28 d of exposure. There was a noticeable decrease after 28 d of exposure (p o0.05) to BP-4, while no significant changes in the BP-2 treatment groups until the end of the experiments (p 40.05). GSH level is shown in Fig. 1D. After 7 d of exposure, GSH levels were significantly induced (p o0.05 or p o0.01) in higher dose groups, BP-1 (5) and BP-3 (5) groups, and remained high and were significantly different from the control values until the end of the experiment, while in lower dose groups significant induced effects were found only in BP-3 (0.5) and BP-4 (0.5) treatment groups after 7 d of exposure, and BP-1 (0.5) treatment group after 7 or 14 d of exposure. However, after 28 d of exposure, GSH level was significantly decreased in BP-2 (5) and all BP-4 treated groups (p o0.05 or p o0.01). 3.2. Integrated biomarker response The transformed data of the studied biomarkers after 28 d of exposure are presented as star plot in Fig. 2A. The calculated IBR values are shown in Fig. 2B. According to this index, the IBR values for the BPs treated groups ranged from 12.8 in BP-3 to 18.9 in BP-1, suggesting that for 28 d exposures, the fish livers in the BP-1 group experienced the most stress, and the least stressful condition was the control. 3.3. Histological alterations To more clearly understand the toxicity of BPs, the histological photomicrographs of liver sections from C. auratus specimens are shown in Fig. 3. After 28 d of exposure, in the normal liver,
0.5mg L-1
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Fig. 2. Biomarker star plots (A) and the calculated IBR values (B) of BPs for different concentrations after 28 d.
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hepatocytes and nuclei were uniform in size and shape (Fig. 3A). Liver specimens exposed to BPs (0.5 mg L 1) showed a slight increase in the hepatocyte damages and these alterations were characterized by the nonuniform distribution and vacuolization of hepatocytes, pyknosis, karyorrhexis, karyolysis (Fig. 3B–E, all graphs were not shown for similarity). Aside from these alterations, clusters of hepatocytes were arranged in a circular wreathlike pattern in the BP-2 and BP-3 groups (Fig. 3C and D).
4. Discussion 4.1. Antioxidant responses
Fig. 3. Representative light micrographs of fish liver after 28 d exposure to 0.5 mg L 1 BPs (hematoxylin–eosin stain, 400 ). (A) Control. (B) BP-1-treated fish, showing hepatic pyknosis (yellow arrow), karyorrhexis (green arrow), and karyolysis (blue arrow). (C) BP-2-treated fish, showing hepatic pyknosis, karyorrhexis and clusters of hepatocytes arranged in a circular wreath-like pattern (rectangle). (D) BP-3-treated fish, showing hepatic pyknosis, karyorrhexis, karyolysis and clusters of hepatocytes arranged in a circular wreath-like pattern. (E) BP-4-treated fish, showing hepatic pyknosis, and karyorrhexis. (For interpretation of the references to color in this figure legend, the reader is referred to the web version of this article.)
In the present study, after 7 d of exposure, SOD and CAT activities were significantly decreased in all treated groups than in the control. This remarkable reduction was presumably due to increasing amounts of ROS (Puerto et al., 2010), which exceeded the removal ROS capacity of SOD and CAT. After 14 or 28 d of exposure, SOD and CAT activities increased gradually and returned to the blank level in most BPs treated groups, suggesting the adaptive response to toxicant stress and a compensatory mechanism to defend against oxidative stress. However, after 28 d of exposure, SOD activity again decreased in the all high-concentration treated groups while CAT activity decreased only in the BP-2 groups, possibly indicating that excessive ROS production resulted in the accumulation of the oxidative substances and antioxidant system failed to scavenge these reactive substances with the prolonged exposure. Besides, after 14 d of exposure, a significant induction of CAT observed in the BP-4 (0.5) group indicated that the generation of H2O2 is still within its elimination capacity. These responses contained the “adaptive stage” and “inhibitive stage” (Li et al., 2012). Generally, the adaptive response to oxidative stress suggested that the antioxidant system attempted to counteract the ROS produced by toxicants; while the inhibitive response of the enzymes under oxidative stress suggested a possible failure of the antioxidant system in clearing H2O2 produced in fish liver and keeping the antioxidant defense balance (Li et al., 2010). Moreover, GST was found to be higher than that in controls in liver of C. auratus exposed to BP-1 and BP-3, suggesting the involvement of this enzyme in the biotransformation of BPs and/or in the antioxidant defense, and the formation of GSH and BPs complexes as a means of detoxification/elimination. Such a mechanism has been previously noted in clams Ruditapes philippinarum exposed to organochlorine pesticide endosulfan (Tao et al., 2013). Although the activation of such a mechanism probably conferred cytoprotection against BPs-induced cellular stress, the toxic effects of BP-1 and BP-3 were apparently not fully neutralized, since changes were observed in other biochemical parameters. GSH levels were instantly induced in the BP-1 (5) and BP-3 (5) groups even in the lower concentration groups, indicating an adaptive and protective role of GSH against oxidative stress induced by chemical contaminants. Zhang et al. studied GSH activity in C. auratus exposed to 2,4-dichlorophenol for 40 days and also found high GSH levels in liver. They suggested that, during moderate oxidative stress, GSH levels in fish liver can increase through increased synthesis as an adaptive mechanism (Zhang et al., 2004). Remarkably, after 28 d of exposure, GSH levels significantly decreased in BP-2 (5) and all BP-4 groups, resulting in the imbalance between the oxidative and antioxidant systems with longer exposure time. As to GSH level, Gao et al. also found a significant decrease in the Tetrahymena thermophile exposed to 1.0 μg L 1 BP-3 (Gao et al., 2013).
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4.2. Comparison of the oxidative stress-inducing capability of four BPs in fish liver In the present study, not all BPs-treated groups displayed significant biochemical response. Thus, the responses triggered by different BPs were apparently different. In order to compare the overall oxidative stress in fish, the IBR values in different groups were calculated and compared. The approach provides a global index of environmental stress by combining the different biomarker signals. Star plots using IBR values instead of biomarker data make it possible to visualize between-site and/or betweensurvey differences for comparison with exposure conditions (Beliaeff and Burgeot, 2002). In this study, IBR values showed that BP1 was most stressful to the fish, followed by BP-2, BP-4 and BP-3. The results may help us to know the oxidative stress of BPs. The IBR, as an indicator of environmental stress, appears to be a useful tool for scientists and managers in assessing ecological risk. However, additional studies should be performed to provide more insight into the mechanism of BPs toxicity in fish. 4.3. Histopathology Microscopic examination of organs has become a vital tool in detecting early effects on morphology (Guardiola et al. 2013a), and has been used as a biomarker to evaluate the toxicity of various contaminants (Velma and Tchounwou, 2010; Guardiola et al., 2013b). In the present study, micrographs of the fish liver exposed to BPs exhibited different degrees of damage, and this could be explained by increased lipid peroxidation and oxidative stress caused by ROS production (Reuter et al. 2010). Two of the main alterations observed in fish were vacuolar degeneration changes and areas of necrosis, which indicates that liver damage occurred in exposed fish. These alterations have been also described in other species of fish living in contaminated environments, suggesting that these alterations might be related to the exposure to environmental chemicals present (Ameur et al. 2012). Vacuolization, which tends to be uniformly distributed, is often especially apparent in the livers of captive fishes and the fishes that are exposed to toxic contaminants. Pycknotic is the state of condensed nuclei present in the hepatocytes, and might be due to the deposition of lipids and glycogen. In the previous reports, the morphology of the testes in adult zebrafish males was not altered after BP-3 treatment (Bluthgen et al., 2012). Dong et al. reported that the yellow catfish fed with the high-cyanobacteria diet showed remarkable histopathological alterations including clusters of hepatocytes arranged in a circular wreath-like pattern (Dong et al., 2012).
5. Conclusions In this study, UV filters including all four BPs tested are generally perceived as safe chemicals suitable for personal care products. Limited data on their biological effects, however, suggested otherwise as the contaminants of emerging concerns. The oxidative stress study showed that the activities of SOD, CAT and GST, the GSH levels and histopathology were altered by BPs exposure, which demonstrated that BPs increase the production of ROS and, thus trigger oxidative stress in liver of C. auratus. Furthermore, the calculated IBR values indicated that different BPs induced different oxidative stress to fish, and the order was ranked as BP-1 4BP2 4BP-44 BP-3. In summary, the parameters measured in this work provide useful baseline information for evaluating the toxicological effects of BPs on fish, but further studies are needed to explore the toxic mechanism of BPs.
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Acknowledgments This research was financially supported by the National Natural Science Foundation of China (Nos. 41071319, and 21377051), the Major Science and Technology Program for Water Pollution Control and Treatment of China (No. 2012ZX07506-001), the Scientific Research Foundation of Graduate School of Nanjing University (2013CL08) and the Planned Science and Technology Project of Jiaxing (2014AY21013).
References Almeida, C., Stepkowska, A., Alegre, A., Nogueira, J.M.F., 2013. Determination of trace levels of benzophenone-type ultra-violet filters in real matrices by bar adsorptive micro-extraction using selective sorbent phases. J. Chromatogr. A 1311, 1–10. Ameur, W.B., de Lapuente, J., Megdiche, Y.E., Barhoumi, B., Trabelsi, S., Camps, L., Serret, J., Ramos-L Ram, D., Gonzalez-Linares, J., Driss, M.R., Borràs, M., 2012. Oxidative stress, genotoxicity and histopathology biomarker responses in mullet (Mugil cephalus) and sea bass (Dicentrarchus labrax) liver from Bizerte Lagoon (Tunisia) Mar. Pollut. Bull. 64, 241–251. Asimakopoulos, A.G., Thomaidis, N.S., Kannan, K., 2014a. Widespread occurrence of bisphenol A diglycidyl ethers, p-hydroxybenzoic acid esters (parabens), benzophenone type-UV filters, triclosan, and triclocarban in human urine from Athens, Greece. Sci. Total Environ. 470–471, 1243–1249. Asimakopoulos, A.G., Wang, L., Thomaidis, N.S., Kannana, K., 2014b. A multi-class bioanalytical methodology for the determination of bisphenol A diglycidyl ethers, p-hydroxybenzoic acid esters, benzophenone-type ultraviolet filters, triclosan, and triclocarban in human urine by liquid chromatography–tandem mass spectrometry. J. Chromatogr. A 1324, 141–148. Balmer, M.E., Buser, H.R., Muller, M.D., Poiger, T., 2005. Occurrence of some organic UV filters in wastewater, in surface waters, and in fish from Swiss lakes. Environ. Sci. Technol. 39, 953–962. Beel, R., Eversloh, C.L., Ternes, T.A., 2013. Biotransformation of the UV-filter sulisobenzone: challenges for the identification of transformation products. Environ. Sci. Technol. 47, 6819–6828. Beliaeff, B., Burgeot, T., 2002. Integrated biomarker response: A useful tool for ecological risk assessment. Environ. Toxicol. Chem. 21, 1316–1322. Bluthgen, N., Zucchi, S., Fent, K., 2012. Effects of the UV filter benzophenone-3 (oxybenzone) at low concentrations in zebrafish (Danio rerio). Toxicol. Appl. Pharmacol. 263, 184–194. Bradford, M.M., 1976. A rapid and sensitive method for quantitation of microgram quantities of protein utilizing the principle of protein-dye-binding. Anal. Biochem. 7, 248–254. Chisvert, A., Leon-Gonzalez, Z., Tarazona, I., Salvador, A., Giokas, D., 2012. An overview of the analytical methods for the determination of organic ultraviolet filters in biological fluids and tissues. Anal. Chim. Acta 752, 11–29. Coronado, M., De Haro, H., Deng, X., Rempel, M.A., Lavado, R., Schlenk, D., 2008. Estrogenic activity and reproductive effects of the UV-filter oxybenzone (2hydroxy-4-methoxyphenylmethanone) in fish. Aquat. Toxicol. 90, 182–187. Dong, G., Xie, S., Zhu, X., Han, D., Yang, Y., Song, L., Gan, L., Chen, W., 2012. Responses of yellow catfish (Pelteobagrus fulvidraco Richardson) exposed to dietary cyanobacteria and subsequent recovery. Toxicon 60, 1298–1306. Downs, C.A., Kramarsky–Winter, E., Fauth, J.E., Segal, R., Bronstein, O., Jeger, R., Lichtenfeld, Y., Woodley, C.M., Pennington, P., Kushmaro, A., Loya, Y., 2014. Toxicological effects of the sunscreen UV filter, benzophenone-2, on planulae and in vitro cells of the coral, Stylophora pistillata. Ecotoxicology 23, 175–191. Duirk, S.E., Bridenstine, D.R., Leslie, D.C., 2013. Reaction of benzophenone UV filters in the presence of aqueous chlorine: kinetics and chloroform formation. Water Res. 47, 579–587. Fent, K., Zenker, A., Rapp, M., 2010. Widespread occurrence of estrogenic UV-filters in aquatic ecosystems in Switzerland. Environ. Pollut. 158, 1817–1824. Gago-Ferrero, P., Alonso, M.B., Bertozzi, C.P., Marigo, J., Barbosa, L., Cremer, M., Secchi, E.R., Azevedo, A., Lailson-Brito Jr., J., Torres, J.P.M., Malm, O., Eljarrat, E., Diaz-Cruz, M.S., Barcelo, D., 2013. First determination of UV filters in marine mammals. Octocrylene levels in franciscana dolphins. Environ. Sci. Technol. 47, 5619–5625. Gao, L., Yuan, T., Zhou, C.Q., Cheng, P., Bai, Q.F., Ao, J.J., Wang, W.H., Zhang, H.M., 2013. Effects of four commonly used UV filters on the growth, cell viability and oxidative stress responses of the Tetrahymena thermophile. Chemosphere 93, 2507–2513. Góth, L., 1991. A simple method for determination of serum catalase activity and revision of reference range. Clin. Chim. Acta 196, 143–151. Grabicova, K., Fedorova, G., Burkina, V., Steinbach, C., Schmidt-Posthaus, H., Zlabek, V., Kroupova, H.K., Grabic, R., Randak, T., 2013. Presence of UV filters in surface water and the effects of phenylbenzimidazole sulfonic acid on rainbow trout (Oncorhynchus mykiss) following a chronic toxicity test. Ecotoxicol. Environ. Saf. 96, 41–47. Guardiola, F.A., Cuesta, A., Meseguer, J., Martinez, S., Martinez-Sanchez, M.J., PerezSirvent, C., Esteban, M.A., 2013a. Accumulation, histopathology and immunotoxicological effects of waterborne cadmium on gilthead seabream
122
H. Liu et al. / Ecotoxicology and Environmental Safety 119 (2015) 116–122
(Sparus aurata). Fish Shellfish Immun. 35, 792–800. Guardiola, F.A., Gonzalez-Parraga, M.P., Cuesta, A., Meseguer, J., Martinez, S., Martinez-Sanchez, M.J., Perez-Sirvent, C., Esteban, M.A., 2013b. Immunotoxicological effects of inorganic arsenic on gilthead seabream (Sparus aurata L.). Aquat. Toxicol. 134–135, 112–119. Han, J., Won, E.J., Hwang, D.S., Rhee, J.S., Kim, I.C., Lee, J.S., 2013. Effect of copper exposure on GST activity and on the expression of four GSTs under oxidative stress condition in the monogonont rotifer, Brachionus koreanus. Comp. Biochem. Phys. C 158, 91–100. Hofkamp, L., Bradley, S., Tresguerres, J., Lichtensteiger, W., Schlumpf, M., Timms, B., 2008. Region-specific growth effects in the developing rat prostate following fetal exposure to estrogenic ultraviolet filters. Environ. Health Perspect. 116, 867–872. Ji, Y.F., Zhou, L., Zhang, Y., Ferronato, C., Brigante, M., Mailhot, G., Yang, X., Chovelon, J.M., 2013. Photochemical degradation of sunscreen agent 2-phenylbenzimidazole-5-sulfonic acid in different water matrices. Water Res. 47, 5865–5875. 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. Jurado, A., Gago-Ferrero, P., Vazquez-Sune, E., Carrera, J., Pujades, E., Diaz-Cruz, M.S., Barcelo, D., 2014. Urban groundwater contamination by residues of UV filters. J. Hazard. Mater. 271, 141–149. Kerdivel, G., Guevel, R.L., Habauzit, D., Brion, F., Ait-Aissa, S., Pakdel, F., 2013. Estrogenic potency of benzophenone UV filters in breast cancer cells: Proliferative and transcriptional activity substantiated by docking analysis. PLoS One 8, 60567. Kunisue, T., Chen, Z., Louis, G.M.B., Sundaram, R., Hediger, M.L., Sun, L., Kannan, K., 2012. Urinary concentrations of benzophenonetype UV filters in U.S. women and their association with endometriosis. Environ. Sci. Technol. 46, 4624–4632. Kunz, P.Y., Galicia, H.F., Fent, K., 2006. Comparison of in vitro and in vivo estrogenic activity of UV filters in fish. Toxicol. Sci. 90, 349–361. Laurentiis, E.D., Minella, M., Sarakha, M., Marrese, A., Minero, C., Mailhot, G., Brigante, M., Vione, D., 2013. Photochemical processes involving the UV absorber benzophenone-4 (2-hydroxy-4-methoxybenzophenone-5-sulphonic acid) in aqueous solution: reaction pathways and implications for surface waters. Water Res. 47, 5943–5953. Li, Y., Li, M., Shi, J.Q., Yang, X., Wang, Z.Y., 2012. Hepatic antioxidative responses to PCDPSs and estimated short-term biotoxicity in freshwater fish. Aquat. Toxicol. 120–121, 90–98. Li, Z.H., Velisek, J., Zlabek, V., Grabic, R., Machova, J., Kolarova, J., Randak, T., 2010. Hepatic antioxidant status and hematological parameters in rainbow trout, Oncorhynchus mykiss, after chronic exposure to carbamazepine. Chem.-Biol. Interact. 183, 98–104. Liao, C.Y., Kannan, K., 2014. Widespread occurrence of benzophenone-type UV light filters in personal care products from China and the United States: an assessment of human exposure. Environ. Sci. Technol. 48, 4103–4109. Liu, H., Sun, P., Liu, H.X., Yang, S.G., Wang, L.S., Wang, Z.Y., 2015. Acute toxicity of benzophenone-type UV filters for Photobacterium phosphoreum and Daphnia magna: QSAR analysis, interspecies relationship and integrated assessment. Chemosphere 135, 182–188. McCord, J.M., Fridovich, I., 1969. Superoxide dismutase an enzymic function for erythrocuprein (hemocuprein). J. Biol. Chem. 244, 6049–6055. Morohoshi, K., Yamamoto, H., Kamata, R., Shiraishi, F., Koda, T., Morita, M., 2005. Estrogenic activity of 37 components of commercial sunscreen lotions evaluated by in vitro assays. Toxicol. In Vitro 19, 457–469. Mottaleb, M.A., Usenko, S., O’Donnell, J.G., Ramirez, A.J., Brooks, B.W., Chambliss, C. K., 2009. Gas chromatography–mass spectrometry screening methods for select UV filters, synthetic musks, alkylphenols, an antimicrobial agent, and an insect repellent in fish. J. Chromatogr. A 1216, 815–823. Paredes, E., Perez, S., Rodil, R., Quintana, J.B., Beiras, R., 2014. Ecotoxicological evaluation of four UV filters using marine organisms from different trophic levels Isochrysis galbana, Mytilus galloprovincialis, Paracentrotus lividus, and Siriella armata. Chemosphere 104, 44–50.
Puerto, M., Pichardo, S., Jos, A., Prieto, A.I., Sevilla, E., Trias, J.E., Camean, A.M., 2010. Differential oxidative stress responses to pure microcystin-LR and microcystincontaining and non-containing cyanobacterial crude extracts on Caco-2 cells. Toxicon 55, 514–522. Qu, R.J., Feng, M.B., Wang, X.H., Qin, L., Wang, C., Wang, Z.Y., Wang, L.S., 2014. Metal accumulation and oxidative stress biomarkers in liver of freshwater fish Carassius auratus following in vivo exposure to waterborne zinc under different pH values. Aquat. Toxicol. 150, 9–16. Reuter, S., Gupta, S.C., Madan, M., Chaturvedi, M.M., Aggarwal, B.B., 2010. Oxidative stress in flammation and cancer: how are they linked? Free Radical Biol. Med. 49, 1603–1616. Tao, Y.X., Pan, L.Q., Zhang, H., Tian, S.M., 2013. Assessment of the toxicity of organochlorine pesticide endosulfan in clams Ruditapes philippinarum. Ecotoxicol. Environ. Saf. 93, 22–30. Tarazona, I., Chisvert, A., Leon, Z., Salvador, A., 2010. Determination of hydroxylated benzophenone UV filters in sea water samples by dispersive liquid–liquid microextraction followed by gas chromatography–mass spectrometry. J. Chromatogr. A 1217, 4771–4778. Tsai, D.Y., Chen, C.L., Ding, W.H., 2014. Optimization of matrix solid-phase dispersion for the rapid determination of salicylate and benzophenone-type UV absorbing substances in marketed fish. Food Chem. 154, 211–216. Tsui, M.M.P., Leung, H.W., Lam, P.K.S., Murphy, M.B., 2014. Seasonal occurrence, removal efficiencies and preliminary risk assessment of multiple classes of organic UV filters in wastewater treatment plants. Water Res. 53, 58–67. Vela-Soria, F., Ballesteros, O., Zafra-Gomez, A., Ballesteros, L., Navalon, A., 2014. A new method for the determination of benzophenone-UV filters in human serum samples by dispersive liquid–liquid microextraction with liquid chromatography–tandem masss pectrometry. Talanta 121, 97–104. Velma, V., Tchounwou, P.B., 2010. Chromium-induced biochemical, genotoxic and histopathologic effects in liver and kidney of goldfish, Carassius auratus. Mutat. Res.—Genet. Toxicol. Environ. Mutagen. 698, 43–51. Wang, L., Asimakopoulos, A.G., Moon, H.B., Nakata, H., Kannan, K., 2013a. Benzotriazole, benzothiazole, and benzophenone compounds in indoor dust from the United States and East Asian countries. Environ. Sci. Technol. 47, 4752–4759. Wang, L., Kannan, K., 2013b. Characteristic profiles of benzonphenone-3 and its derivatives in urine of children and adults from the United States and China. Environ. Sci. Technol. 47, 12532–12538. Weisbrod, C.J., Kunz, P.Y., Zenker, A.K., Fent, K., 2007. Effects of the UV filter benzophenone-2 on reproduction in fish. Toxicol. Appl. Pharmacol. 225, 255–266. Wu, J.W., Chen, H.C., Ding, W.H., 2013. Ultrasound-assisted dispersive liquid–liquid microextraction plus simultaneous silylation for rapid determination of salicylate and benzophenone-type ultraviolet filters in aqueous samples. J. Chromatogr. A 1302, 20–27. Xiao, M., Wei, D.B., Yin, J.X., Wei, G.H., Du, Y.G., 2013. Transformation mechanism of benzophenone-4 in free chlorine promoted chlorination disinfection. Water Res. 47, 6223–6233. Yang, B., Ying, G.G., 2013. Oxidation of benzophenone-3 during water treatment with ferrate (VI). Water Res. 47, 2458–2466. Zhang, J.F., Shen, H., Wang, X.R., Wu, J.C., Xue, Y.Q., 2004. Effects of chronic exposure of 2,4-dichlorophenol on the antioxidant system in liver of freshwater fish Carassius auratus. Chemosphere 55, 167–174. Zhang, T., Sun, H.W., Qin, X.L., Wu, Q., Zhang, Y.F., Ma, J., Kannan, K., 2013. Benzophenone–type UV filters in urine and blood from children, adults, and pregnant women in China: partitioning between blood and urine as well as maternal and fetal cord blood. Sci. Total Environ. 461–462, 49–55. Zhang, Z.F., Ren, N.Q., Li, Y.F., Kunisue, T., Gao, D.W., Kannan, K., 2011. Determination of benzotriazole and benzophenone UV Filters in sediment and sewage sludge. Environ. Sci. Technol. 45, 3909–3916. Zucchi, S., Bluthgen, N., Ieronimo, A., 2011. The UV-absorder benzophenone-4 alters transcripts of genes involved in hormonal pathways in zebrafish (Danio rerio) eleuthero-embryo s and adult males. Toxicol. Appl. Pharmacol 250, 137–146.
Contents lists available at ScienceDirect
Ecotoxicology and Environmental Safety journal homepage: www.elsevier.com/locate/ecoenv
Hepatic oxidative stress biomarker responses in freshwater fish Carassius auratus exposed to four benzophenone UV filters Hui Liu a,b, Ping Sun b, Hongxia Liu b, Shaogui Yang a, Liansheng Wang a, Zunyao Wang a,n a b
State Key Laboratory of Pollution Control and Resources Reuse, School of Environment, Nanjing University, Jiangsu Nanjing 210023, PR China College of Biological and Chemical Engineering, Jiaxing University, Zhejiang Jiaxing 314001, PR China
art ic l e i nf o
a b s t r a c t s
Article history: Received 15 October 2014 Received in revised form 10 May 2015 Accepted 11 May 2015
Benzophenone (BP) type UV filters are widely used in many personal care products to protect human from UV irradiation. Despite the estrogenic potencies to fish and the environmental occurrences of BP derivatives in aquatic systems, little information is available regarding their effects on the antioxidant defense system in fish. In this work, the oxidative stress induced 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. Moreover, liver tissues were also studied histologically. The changes in the activities of antioxidant enzymes and glutathione levels suggested that BPs generates oxidative stress in fish. The IBR index revealed that the hepatic oxidative toxicity followed the order BP-1 4BP-2 4BP4 4BP-3. The histopathological analysis revealed lesions caused by BPs. This investigation provides essential information for assessing the potential ecological risk of BP-type UV filters. & 2015 Elsevier Inc. All rights reserved.
Keywords: UV filters Benzophenones Oxidative stress Integrated biomarker response Histological analysis
1. Introduction Benzophenones (BPs) are one of the primary components in the UV filters family and are, extensively used as sunscreen agents in personal care products (PCPs) for the protection of skin and hair from UV irradiation (Chisvert et al., 2012). In the European Union, 2-hydroxyl-4-methoxyl benzophenone (BP-3) and 2-hydroxyl-4methoxyl benzophenone-5-sulfonic acid (BP-4) are approved to be used as UV filters in sunscreens at maximum individual concentrations of 10% and 5%, respectively, and in Japan, 2,4-dihydroxybenzophenone (BP-1), 2,2′,4,4′-tetrahydroxybenzophenone (BP-2) and 2,2′-dihydroxy-4,4′-dimethoxybenzophenone (BP-6) are also allowed in sunscreens. Currently, BPs have been classified as chemicals suspected of having endocrine disrupting effects due to increasing evidence that BPs are able to interfere with the endocrine system (Kerdivel et al., 2013). BPs have been detected in environmental matrices such as water (Almeida et al., 2013; Wu et al., 2013; Jurado et al., 2014), soil, sediment (Zhang et al., 2011) and indoor dust (Wang et al., 2013a). The amounts found in water samples are in the ng L 1 range, which are not far below the doses that cause toxic effects in animals (Tarazona et al., 2010). Fish are important organisms to monitor the occurrence of persistent lipophilic contaminants (Fent et al., 2010). The presence of BPs UV filters was reported in fish n
Corresponding author. Fax: þ86 25 89680358. E-mail address: [email protected] (Z. Wang).
http://dx.doi.org/10.1016/j.ecoenv.2015.05.017 0147-6513/& 2015 Elsevier Inc. All rights reserved.
from Switzerland rivers, Swiss lakes and Texas streams (Balmer et al., 2005; Mottaleb et al., 2009; Fent et al., 2010), even in marketed fish (Tsai et al., 2014). Some BP UV-filters have hormonal activity in fish. It is reported that BP-1, BP-2, BP-3 and BP-4 elicit estrogenic and anti-androgenic activities in fish (Kunz et al., 2006) in vivo and in vitro tests, and it is noteworthy that BP-1, BP-2 and 4-hydroxybenzophenone (4BP) possess estrogenic activities higher than that of BP-3 (Morohoshi et al., 2005). BP-2 could negatively affect egg production and gonadal development in fish (Weisbrod et al., 2007). In addition, induction of vitellogenin and impairment of reproduction in fish (Japanese medaka and rainbow trout), and alterations of gene expression in both adult zebrafish and eleuthero-embryos were reported with an aqueous exposure to BP-3 (Coronado et al., 2008; Bluthgen et al., 2012). Alterations of gene expression involved in hormonal pathways and steroidogenesis were also reported in zebrafish exposed to BP-4 (Zucchi et al., 2011). BPs have even been detected in human samples, such as breast milk, urine (Asimakopoulos et al., 2014a, 2014b) and blood (Zhang et al., 2013; Vela-Soria et al., 2014). Recently, through concentration analysis of BP-3 as well as four of its metabolic derivatives, BP1, BP-2, 2,2′-dihydroxyl-4-methoxl benzophenone (BP-8), and 4BP in urine of children and adults, BP-3 was found in nearly all urine samples from the U.S. and China (Wang and Kannan, 2013b). High urinary concentrations of BP-1 are associated with estrogen-dependent diseases, such as endometriosis in women (Kunisue et al., 2012). These findings demonstrate the apparent likelihood that BPs contained in PCPs might be emitted into the aquatic
H. Liu et al. / Ecotoxicology and Environmental Safety 119 (2015) 116–122
environment and persistent with a great potential for bioaccumulation through food chains (Gago-Ferrero et al., 2013). Thus, the occurrence (Liao and Kannan, 2014; Tsui et al., 2014), the degradation (Beel et al., 2013; Duirk et al., 2013; Yang and Ying, 2013; Ji et al., 2013; Laurentiis et al., 2013; Xiao et al., 2013) and, particularly the potential toxicities to humans and the ecosytems of these BP UV-filters have attracted widespread attention (Hofkamp et al., 2008; Downs et al., 2014; Paredes et al., 2014; Liu et al., 2015). Biomarkers such as enzyme activity, protein level and DNA can be used to measure the interaction between biological systems and physical, chemicals or biological environmental agents. And the evaluation of oxidative stress biomarkers is critical to the investigation of oxidative stress in organisms. Grabicova et al. found, after 42 days of UV filter 2-phenylbenzimidazole-5-sulfonicacid (PBSA) exposure in rainbow trout liver, the catalase (CAT) and superoxide dismutase (SOD) activities were not significantly different between the control group and the group that was exposed to the highest concentration of PBSA (1000 mg L 1), and no pathological changes were obvious in the liver or gonads (Grabicova et al., 2013). Gao et al. observed increased CAT activity in the Tetrahymena thermophile exposed to 1.0 μg L 1 BP-3, while no significant changes of the SOD activity (Gao et al., 2013). Therefore, we attempted to test the hypothesis whether these environmentally important benzophenones induce oxidative stress in fish. In the present study, four oxidation-related biomarkers, the activities of SOD, CAT and glutathione S-transferase (GST), and the level of glutathione (GSH) were measured to estimate the single toxicities of four common BP-type UV filters on antioxidant status in fish Carassius auratus (C. auratus). Next, integrated biomarker response (IBR) index was applied to improve their discriminatory power, and was calculated to estimate the integral effects of BPsinduced oxidative stress. Finally, liver tissues were examined histologically.
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Table 1 Water quality parameters of the water used for acclimation and subsequent experiments. Water quality parameters pH Conductivity Total hardness Alkalinity Na þ
7.25 7 0.25 350.67 12.5 μs/cm 146.5 7 9.3 mg CaCO3/L 40.5 75.0 mg CaCO3/L 12.2 70.2 mg/L
Kþ Mg2 þ Ca2 þ Cl DO (dissolved oxygen)
2.34 7 0.05 mg/L 7.75 7 0.02 mg/L 42.05 7 0.82 mg/L 28.17 1.2 mg/L 6.5 7 0.5 mg O2/L
water. The solubility of BP-3 is the most smallest than others, and the saturated solubility of BP-3 is 5.1 mg L 1 (Gago-Ferrero et al., 2013). Thus the lower concentration 0.5 mg L 1 were chose as close to worst-case environmental concentrations, the higher one of 5 mg L 1 as a pharmacological concentration. BPs were analyzed on an Agilent 1200 High Performance Liquid Chromatograph (HPLC) equipped with a Diode Array Detector. A Zorbax 300 SBC18 column (4.6 150 mm, 5 μm) was used for the separation at 30 °C. The mobile phase was 0.3% formic acid in water (A) and methanol (B) with an isocratic elution of 20:80 (v/v). The injection volume was 100 mL, the flow rate was set at 1 mL min 1, and the detection wavelength was 290 nm. The actual exposure concentrations of BP-1, BP-2, BP-3 and BP-4 in low dose groups are 0.48 70.03, 0.487 0.05, 0.48 70.03 and 0.497 0.03 mg L 1, respectively, and in high dose groups 4.707 0.47, 4.78 70.40, 4.767 0.32 and 4.90 70.40 mg L 1, respectively. Every two days, a half of water in the tanks was exchanged with fresh water containing the same concentrations of BPs to maintain the constant concentrations during the experiments. The final DMSO concentration was 0.005% in all tanks except the aqueous control. The protocols for fish maintaining, experimentation and sacrifice were approved by the Ethics Committee of Nanjing University (Qu et al., 2014).
2. Materials and methods
2.4. Sample preparation and biochemical assessments
2.1. Chemicals and reagents
4 fish were randomly sampled from each group on days 7, 14 and 28, and then killed by a blow to the head and dissected for liver in each treatment. Subsequently, the liver of each fish was placed on an ice plate and rinsed with cold physiological saline (0.9% NaCl), next weighed, and homogenized with cold physiological saline using an Ultra Turrax homogenizer (IKA, Germany). The homogenates were centrifuged by Eppendorf 5417R centrifuge (Eppendorf, Germany) at 4000 rpm at 4 °C for 15 min, to get the supernatants for further biochemical analysis. Four biomarkers, including the activities of SOD, CAT and GST, and GSH level, were analyzed. The SOD activity was measured based on the inhibition of cytochrome c caused by the superoxide radical (McCord and Fridovich, 1969). The CAT activity was evaluated by monitoring residual H2O2 absorbance at 405 nm (Góth, 1991). The GST activity was measured using GST assay kit by a spectrophotometer with 1-chloro-2,4-dinitrobenzene (CDNB) and glutathione-reduced (GSH) as substrates at 412 nm (Han et al., 2013). The GSH level was evaluated following the procedure of Jollow et al. (1974) with 5,5′dithiobis-2-nitrobenzoic acid (DTNB) as the reagent. DTNB was reduced by free sulfhydryl groups of GSH to form yellow-colored 5-thio-nitrobenzoic acid. The activities of antioxidant enzymes were normalized by total protein and represented as a relative percentage of the control. Total proteins were determined using the Bradford method (Bradford, 1976).
BP-1 (CAS no:131-56-6, 99% purity), BP-2 (CAS no:131-55-5, 99% purity), BP-3 (CAS no:131-57-7, 99% purity) and BP-4 (CAS no:4065-45-6, 98% purity) were purchased from J&K Co. (Shanghai, China). The stock solutions were prepared in dimethylsulfoxide (DMSO). All other reagents were of analytical purity. The commercial assay kits for the analysis of the oxidative stress biomarkers were purchased from Nanjing Jiancheng Bioengineering Institute (Nanjing, China). 2.2. 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. 2.3. Animals and experimental exposure C. auratus (weight: 28.84 73.35 g) were obtained commercially from a local supplier (Nanjing, China). Before the experiment, the healthy fish were acclimatized for at least 7 d in tanks containing dechlorinated and aerated tap water at 2472 °C. The fish were fed with the commercial fish food twice a day but fasted 24 h prior to biochemical analysis. During the acclimated period, the total mortality was less than 1%. After acclimatization, fifteen acclimated fish were transferred to each aquarium containing 30 L dechlorinated and aerated
2.5. Integrated biomarker response (IBR) A battery of biomarkers is often used to evaluate the effects of
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exposure to chemical contaminants and detect responses to environmental stress. Unfortunately, field application of biomarkers is subject to various constraints that can limit data acquisition and prevent the use of multivariate methods during statistical analysis. Therefore, a simple method is needed to summarize biomarker responses and simplify their interpretation in biomonitoring programs. Integrated biomarker response is a method to summarize the biomarker responses into one general “stress index”. It was applied to assess the potential toxicity of BPs to fish. The procedures for the IBR calculation are described here briefly: (1) data were calculated for mean and SD (standard deviation). (2) Data were standardized according to the formula Y ¼(X m)/S, where Y is the standardized data, X is the data of each biomarker response, m is the mean data of the biomarker, and S is the standard deviation of the biomarker. (3) Z was calculated as Z¼Y in the case of activation or Z¼ Y in the case of inhibition. The minimum value (Min) was obtained. (4) S was calculated as S ¼Zþ ❘Min❘, where S Z0 and ❘Min❘ is the absolute value. (5) Calculation of star plot areas by multiplying the obtained value of each biomarker (Si) with the next biomarker (Si þ 1) and dividing each calculation by 2. (6) The corresponding IBR value was obtained as
80 *
70
Histopathological examination of fish liver was performed using light microscopy. The liver was sampled, immobilized in 10% formaldehyde solution, wrapped in wax, and cut into slices of 4– 5 μm thickness using an ultra-thin semiautomatic microtome (ERM-3000, China). The slices were later stained with hematoxylin and counter stained with eosin. Next, the liver slices were examined under a light microscope (ZEISS Axio Imager A1, Germany) at a magnification of 400 and evaluated for changes in morphology of BPs exposed tissues compared to the aqueous control. 2.7. Statistical analysis Statistical analyses were conducted using the SPSS statistical package (ver. 16.0, SPSS Company, Chicago, USA). Whether the data distributions were normal was determined by the Shapiro–Wilk test and the homogeneity of variances was determined by the
**
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** **
GST (U/mg Protein)
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40
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20
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40 30 20
0 Aq ue ou sC on So tro lv l en tC on tro l BP -1 (0 .5 ) BP -1 (5 ) BP -2 (0 .5 ) BP -2 (5 ) BP -3 (0 .5 ) BP -3 (5 ) BP -4 (0 .5 ) BP -4 (5 )
0
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90
7d 14d 28d
80
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GSH (mmol/mg Protein)
CAT activity (U/mg Protein)
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50
10
20
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60
10
10
**
70
*
** **
**
90
Aq ue ou sC on So tro lv l en tC on tro l BP -1 (0 .5 ) BP -1 (5 ) BP -2 (0 .5 ) BP -2 (5 ) BP -3 (0 .5 ) BP -3 (5 ) BP -4 (0 .5 ) BP -4 (5 )
SOD activity (U/mg Protein)
2.6. Histopathology
7d 14d 28d
90
60
IBR¼ {[(S1 S2)/2] þ[(S2 S3)/2] þ ⋯ þ[(Si Si þ 1)/2] }, when i¼ n, Si þ 1 ¼S1 (Beliaeff and Burgeot, 2002).
60 **
50
30 20
** **
**
40
**
* *
*
*
*
*
10 0 Aq ue ou sC on So tro lv l en tC on tro l BP -1 (0 .5 ) BP -1 (5 ) BP -2 (0 .5 ) BP -2 (5 ) BP -3 (0 .5 ) BP -3 (5 ) BP -4 (0 .5 ) BP -4 (5 )
Aq ue ou sC on So tro lv l en tC on tro l BP -1 (0 .5 ) BP -1 (5 ) BP -2 (0 .5 ) BP -2 (5 ) BP -3 (0 .5 ) BP -3 (5 ) BP -4 (0 .5 ) BP -4 (5 )
0
**
Fig. 1. The effects of BPs on the antioxidant enzymes activities (SOD, CAT and GST) and GSH level in liver of C. auratus. Data are means 7 SD, n ¼4 for each data point. *Different from control (p o 0.05). **Different from control (p o 0.01).
H. Liu et al. / Ecotoxicology and Environmental Safety 119 (2015) 116–122
Levene test. One-way analysis of variance (ANOVA) with the Dunnett’s test was used to analyze differences between the control and the experimental groups. The significant difference contains two grades, significant (p o0.05) and highly significant (p o0.01). Results are expressed as means 7 SD.
3. Results 3.1. Oxidative stress in fish No mortality was observed during the acclimatization and the exposure period to BPs, and there was no significant difference (p 40.05) among the blank and the DMSO control for all biomarkers during the exposure. Changes in the measured oxidative stress biomarkers, including changes in the activities of the antioxidant enzymes activities (SOD, CAT and GST) and changes in the level of the non-enzyme antioxidantare (GSH) are shown in Fig. 1A–D. After 7 d of exposure, there was a significantly decreasing trend in the SOD activity in all BPs-treated groups (p o0.01). After 14 or 28 d of exposure, the SOD activity increased gradually and returned to the blank level, and no significant changes were observed in SOD activity in the lower dose treated groups, except in the high concentration treated groups (Fig. 1A). Similarly, after 7 d of exposure, there was a noteworthy decreasing in CAT activity (p o0.05 or p o0.01) in all dose groups (Fig. 1B). With an increasing expose time, the CAT activity also returned to the blank level gradually, except in all BP-2 treated groups. After 14 or 28 d of exposure, there was a noteworthy decreasing trend (p o0.05 or p o0.01) in the BP-2 group (0.5), and after 28 d of exposure in the BP-2 group (5). Moreover, the CAT activity was induced significantly after 14 d of exposure to BP-4 (0.5).
Fig. 1C shows the activity of GST during the exposure. GST activity increased by BP-1 and BP-3 significantly (po 0.05 or po 0.01) at first and peaked after 7 d of exposure, then remained after 14 d of exposure, but returned to the blank level after 28 d of exposure. There was a noticeable decrease after 28 d of exposure (p o0.05) to BP-4, while no significant changes in the BP-2 treatment groups until the end of the experiments (p 40.05). GSH level is shown in Fig. 1D. After 7 d of exposure, GSH levels were significantly induced (p o0.05 or p o0.01) in higher dose groups, BP-1 (5) and BP-3 (5) groups, and remained high and were significantly different from the control values until the end of the experiment, while in lower dose groups significant induced effects were found only in BP-3 (0.5) and BP-4 (0.5) treatment groups after 7 d of exposure, and BP-1 (0.5) treatment group after 7 or 14 d of exposure. However, after 28 d of exposure, GSH level was significantly decreased in BP-2 (5) and all BP-4 treated groups (p o0.05 or p o0.01). 3.2. Integrated biomarker response The transformed data of the studied biomarkers after 28 d of exposure are presented as star plot in Fig. 2A. The calculated IBR values are shown in Fig. 2B. According to this index, the IBR values for the BPs treated groups ranged from 12.8 in BP-3 to 18.9 in BP-1, suggesting that for 28 d exposures, the fish livers in the BP-1 group experienced the most stress, and the least stressful condition was the control. 3.3. Histological alterations To more clearly understand the toxicity of BPs, the histological photomicrographs of liver sections from C. auratus specimens are shown in Fig. 3. After 28 d of exposure, in the normal liver,
0.5mg L-1
5mg/L 0.5mg/L
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IBR
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BP -4
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tr on nt C ve
So l
qu
eo
us
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on
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ol
ol
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A
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Fig. 2. Biomarker star plots (A) and the calculated IBR values (B) of BPs for different concentrations after 28 d.
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hepatocytes and nuclei were uniform in size and shape (Fig. 3A). Liver specimens exposed to BPs (0.5 mg L 1) showed a slight increase in the hepatocyte damages and these alterations were characterized by the nonuniform distribution and vacuolization of hepatocytes, pyknosis, karyorrhexis, karyolysis (Fig. 3B–E, all graphs were not shown for similarity). Aside from these alterations, clusters of hepatocytes were arranged in a circular wreathlike pattern in the BP-2 and BP-3 groups (Fig. 3C and D).
4. Discussion 4.1. Antioxidant responses
Fig. 3. Representative light micrographs of fish liver after 28 d exposure to 0.5 mg L 1 BPs (hematoxylin–eosin stain, 400 ). (A) Control. (B) BP-1-treated fish, showing hepatic pyknosis (yellow arrow), karyorrhexis (green arrow), and karyolysis (blue arrow). (C) BP-2-treated fish, showing hepatic pyknosis, karyorrhexis and clusters of hepatocytes arranged in a circular wreath-like pattern (rectangle). (D) BP-3-treated fish, showing hepatic pyknosis, karyorrhexis, karyolysis and clusters of hepatocytes arranged in a circular wreath-like pattern. (E) BP-4-treated fish, showing hepatic pyknosis, and karyorrhexis. (For interpretation of the references to color in this figure legend, the reader is referred to the web version of this article.)
In the present study, after 7 d of exposure, SOD and CAT activities were significantly decreased in all treated groups than in the control. This remarkable reduction was presumably due to increasing amounts of ROS (Puerto et al., 2010), which exceeded the removal ROS capacity of SOD and CAT. After 14 or 28 d of exposure, SOD and CAT activities increased gradually and returned to the blank level in most BPs treated groups, suggesting the adaptive response to toxicant stress and a compensatory mechanism to defend against oxidative stress. However, after 28 d of exposure, SOD activity again decreased in the all high-concentration treated groups while CAT activity decreased only in the BP-2 groups, possibly indicating that excessive ROS production resulted in the accumulation of the oxidative substances and antioxidant system failed to scavenge these reactive substances with the prolonged exposure. Besides, after 14 d of exposure, a significant induction of CAT observed in the BP-4 (0.5) group indicated that the generation of H2O2 is still within its elimination capacity. These responses contained the “adaptive stage” and “inhibitive stage” (Li et al., 2012). Generally, the adaptive response to oxidative stress suggested that the antioxidant system attempted to counteract the ROS produced by toxicants; while the inhibitive response of the enzymes under oxidative stress suggested a possible failure of the antioxidant system in clearing H2O2 produced in fish liver and keeping the antioxidant defense balance (Li et al., 2010). Moreover, GST was found to be higher than that in controls in liver of C. auratus exposed to BP-1 and BP-3, suggesting the involvement of this enzyme in the biotransformation of BPs and/or in the antioxidant defense, and the formation of GSH and BPs complexes as a means of detoxification/elimination. Such a mechanism has been previously noted in clams Ruditapes philippinarum exposed to organochlorine pesticide endosulfan (Tao et al., 2013). Although the activation of such a mechanism probably conferred cytoprotection against BPs-induced cellular stress, the toxic effects of BP-1 and BP-3 were apparently not fully neutralized, since changes were observed in other biochemical parameters. GSH levels were instantly induced in the BP-1 (5) and BP-3 (5) groups even in the lower concentration groups, indicating an adaptive and protective role of GSH against oxidative stress induced by chemical contaminants. Zhang et al. studied GSH activity in C. auratus exposed to 2,4-dichlorophenol for 40 days and also found high GSH levels in liver. They suggested that, during moderate oxidative stress, GSH levels in fish liver can increase through increased synthesis as an adaptive mechanism (Zhang et al., 2004). Remarkably, after 28 d of exposure, GSH levels significantly decreased in BP-2 (5) and all BP-4 groups, resulting in the imbalance between the oxidative and antioxidant systems with longer exposure time. As to GSH level, Gao et al. also found a significant decrease in the Tetrahymena thermophile exposed to 1.0 μg L 1 BP-3 (Gao et al., 2013).
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4.2. Comparison of the oxidative stress-inducing capability of four BPs in fish liver In the present study, not all BPs-treated groups displayed significant biochemical response. Thus, the responses triggered by different BPs were apparently different. In order to compare the overall oxidative stress in fish, the IBR values in different groups were calculated and compared. The approach provides a global index of environmental stress by combining the different biomarker signals. Star plots using IBR values instead of biomarker data make it possible to visualize between-site and/or betweensurvey differences for comparison with exposure conditions (Beliaeff and Burgeot, 2002). In this study, IBR values showed that BP1 was most stressful to the fish, followed by BP-2, BP-4 and BP-3. The results may help us to know the oxidative stress of BPs. The IBR, as an indicator of environmental stress, appears to be a useful tool for scientists and managers in assessing ecological risk. However, additional studies should be performed to provide more insight into the mechanism of BPs toxicity in fish. 4.3. Histopathology Microscopic examination of organs has become a vital tool in detecting early effects on morphology (Guardiola et al. 2013a), and has been used as a biomarker to evaluate the toxicity of various contaminants (Velma and Tchounwou, 2010; Guardiola et al., 2013b). In the present study, micrographs of the fish liver exposed to BPs exhibited different degrees of damage, and this could be explained by increased lipid peroxidation and oxidative stress caused by ROS production (Reuter et al. 2010). Two of the main alterations observed in fish were vacuolar degeneration changes and areas of necrosis, which indicates that liver damage occurred in exposed fish. These alterations have been also described in other species of fish living in contaminated environments, suggesting that these alterations might be related to the exposure to environmental chemicals present (Ameur et al. 2012). Vacuolization, which tends to be uniformly distributed, is often especially apparent in the livers of captive fishes and the fishes that are exposed to toxic contaminants. Pycknotic is the state of condensed nuclei present in the hepatocytes, and might be due to the deposition of lipids and glycogen. In the previous reports, the morphology of the testes in adult zebrafish males was not altered after BP-3 treatment (Bluthgen et al., 2012). Dong et al. reported that the yellow catfish fed with the high-cyanobacteria diet showed remarkable histopathological alterations including clusters of hepatocytes arranged in a circular wreath-like pattern (Dong et al., 2012).
5. Conclusions In this study, UV filters including all four BPs tested are generally perceived as safe chemicals suitable for personal care products. Limited data on their biological effects, however, suggested otherwise as the contaminants of emerging concerns. The oxidative stress study showed that the activities of SOD, CAT and GST, the GSH levels and histopathology were altered by BPs exposure, which demonstrated that BPs increase the production of ROS and, thus trigger oxidative stress in liver of C. auratus. Furthermore, the calculated IBR values indicated that different BPs induced different oxidative stress to fish, and the order was ranked as BP-1 4BP2 4BP-44 BP-3. In summary, the parameters measured in this work provide useful baseline information for evaluating the toxicological effects of BPs on fish, but further studies are needed to explore the toxic mechanism of BPs.
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Acknowledgments This research was financially supported by the National Natural Science Foundation of China (Nos. 41071319, and 21377051), the Major Science and Technology Program for Water Pollution Control and Treatment of China (No. 2012ZX07506-001), the Scientific Research Foundation of Graduate School of Nanjing University (2013CL08) and the Planned Science and Technology Project of Jiaxing (2014AY21013).
References Almeida, C., Stepkowska, A., Alegre, A., Nogueira, J.M.F., 2013. Determination of trace levels of benzophenone-type ultra-violet filters in real matrices by bar adsorptive micro-extraction using selective sorbent phases. J. Chromatogr. A 1311, 1–10. Ameur, W.B., de Lapuente, J., Megdiche, Y.E., Barhoumi, B., Trabelsi, S., Camps, L., Serret, J., Ramos-L Ram, D., Gonzalez-Linares, J., Driss, M.R., Borràs, M., 2012. Oxidative stress, genotoxicity and histopathology biomarker responses in mullet (Mugil cephalus) and sea bass (Dicentrarchus labrax) liver from Bizerte Lagoon (Tunisia) Mar. Pollut. Bull. 64, 241–251. Asimakopoulos, A.G., Thomaidis, N.S., Kannan, K., 2014a. Widespread occurrence of bisphenol A diglycidyl ethers, p-hydroxybenzoic acid esters (parabens), benzophenone type-UV filters, triclosan, and triclocarban in human urine from Athens, Greece. Sci. Total Environ. 470–471, 1243–1249. Asimakopoulos, A.G., Wang, L., Thomaidis, N.S., Kannana, K., 2014b. A multi-class bioanalytical methodology for the determination of bisphenol A diglycidyl ethers, p-hydroxybenzoic acid esters, benzophenone-type ultraviolet filters, triclosan, and triclocarban in human urine by liquid chromatography–tandem mass spectrometry. J. Chromatogr. A 1324, 141–148. Balmer, M.E., Buser, H.R., Muller, M.D., Poiger, T., 2005. Occurrence of some organic UV filters in wastewater, in surface waters, and in fish from Swiss lakes. Environ. Sci. Technol. 39, 953–962. Beel, R., Eversloh, C.L., Ternes, T.A., 2013. Biotransformation of the UV-filter sulisobenzone: challenges for the identification of transformation products. Environ. Sci. Technol. 47, 6819–6828. Beliaeff, B., Burgeot, T., 2002. Integrated biomarker response: A useful tool for ecological risk assessment. Environ. Toxicol. Chem. 21, 1316–1322. Bluthgen, N., Zucchi, S., Fent, K., 2012. Effects of the UV filter benzophenone-3 (oxybenzone) at low concentrations in zebrafish (Danio rerio). Toxicol. Appl. Pharmacol. 263, 184–194. Bradford, M.M., 1976. A rapid and sensitive method for quantitation of microgram quantities of protein utilizing the principle of protein-dye-binding. Anal. Biochem. 7, 248–254. Chisvert, A., Leon-Gonzalez, Z., Tarazona, I., Salvador, A., Giokas, D., 2012. An overview of the analytical methods for the determination of organic ultraviolet filters in biological fluids and tissues. Anal. Chim. Acta 752, 11–29. Coronado, M., De Haro, H., Deng, X., Rempel, M.A., Lavado, R., Schlenk, D., 2008. Estrogenic activity and reproductive effects of the UV-filter oxybenzone (2hydroxy-4-methoxyphenylmethanone) in fish. Aquat. Toxicol. 90, 182–187. Dong, G., Xie, S., Zhu, X., Han, D., Yang, Y., Song, L., Gan, L., Chen, W., 2012. Responses of yellow catfish (Pelteobagrus fulvidraco Richardson) exposed to dietary cyanobacteria and subsequent recovery. Toxicon 60, 1298–1306. Downs, C.A., Kramarsky–Winter, E., Fauth, J.E., Segal, R., Bronstein, O., Jeger, R., Lichtenfeld, Y., Woodley, C.M., Pennington, P., Kushmaro, A., Loya, Y., 2014. Toxicological effects of the sunscreen UV filter, benzophenone-2, on planulae and in vitro cells of the coral, Stylophora pistillata. Ecotoxicology 23, 175–191. Duirk, S.E., Bridenstine, D.R., Leslie, D.C., 2013. Reaction of benzophenone UV filters in the presence of aqueous chlorine: kinetics and chloroform formation. Water Res. 47, 579–587. Fent, K., Zenker, A., Rapp, M., 2010. Widespread occurrence of estrogenic UV-filters in aquatic ecosystems in Switzerland. Environ. Pollut. 158, 1817–1824. Gago-Ferrero, P., Alonso, M.B., Bertozzi, C.P., Marigo, J., Barbosa, L., Cremer, M., Secchi, E.R., Azevedo, A., Lailson-Brito Jr., J., Torres, J.P.M., Malm, O., Eljarrat, E., Diaz-Cruz, M.S., Barcelo, D., 2013. First determination of UV filters in marine mammals. Octocrylene levels in franciscana dolphins. Environ. Sci. Technol. 47, 5619–5625. Gao, L., Yuan, T., Zhou, C.Q., Cheng, P., Bai, Q.F., Ao, J.J., Wang, W.H., Zhang, H.M., 2013. Effects of four commonly used UV filters on the growth, cell viability and oxidative stress responses of the Tetrahymena thermophile. Chemosphere 93, 2507–2513. Góth, L., 1991. A simple method for determination of serum catalase activity and revision of reference range. Clin. Chim. Acta 196, 143–151. Grabicova, K., Fedorova, G., Burkina, V., Steinbach, C., Schmidt-Posthaus, H., Zlabek, V., Kroupova, H.K., Grabic, R., Randak, T., 2013. Presence of UV filters in surface water and the effects of phenylbenzimidazole sulfonic acid on rainbow trout (Oncorhynchus mykiss) following a chronic toxicity test. Ecotoxicol. Environ. Saf. 96, 41–47. Guardiola, F.A., Cuesta, A., Meseguer, J., Martinez, S., Martinez-Sanchez, M.J., PerezSirvent, C., Esteban, M.A., 2013a. Accumulation, histopathology and immunotoxicological effects of waterborne cadmium on gilthead seabream
122
H. Liu et al. / Ecotoxicology and Environmental Safety 119 (2015) 116–122
(Sparus aurata). Fish Shellfish Immun. 35, 792–800. Guardiola, F.A., Gonzalez-Parraga, M.P., Cuesta, A., Meseguer, J., Martinez, S., Martinez-Sanchez, M.J., Perez-Sirvent, C., Esteban, M.A., 2013b. Immunotoxicological effects of inorganic arsenic on gilthead seabream (Sparus aurata L.). Aquat. Toxicol. 134–135, 112–119. Han, J., Won, E.J., Hwang, D.S., Rhee, J.S., Kim, I.C., Lee, J.S., 2013. Effect of copper exposure on GST activity and on the expression of four GSTs under oxidative stress condition in the monogonont rotifer, Brachionus koreanus. Comp. Biochem. Phys. C 158, 91–100. Hofkamp, L., Bradley, S., Tresguerres, J., Lichtensteiger, W., Schlumpf, M., Timms, B., 2008. Region-specific growth effects in the developing rat prostate following fetal exposure to estrogenic ultraviolet filters. Environ. Health Perspect. 116, 867–872. Ji, Y.F., Zhou, L., Zhang, Y., Ferronato, C., Brigante, M., Mailhot, G., Yang, X., Chovelon, J.M., 2013. Photochemical degradation of sunscreen agent 2-phenylbenzimidazole-5-sulfonic acid in different water matrices. Water Res. 47, 5865–5875. 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. Jurado, A., Gago-Ferrero, P., Vazquez-Sune, E., Carrera, J., Pujades, E., Diaz-Cruz, M.S., Barcelo, D., 2014. Urban groundwater contamination by residues of UV filters. J. Hazard. Mater. 271, 141–149. Kerdivel, G., Guevel, R.L., Habauzit, D., Brion, F., Ait-Aissa, S., Pakdel, F., 2013. Estrogenic potency of benzophenone UV filters in breast cancer cells: Proliferative and transcriptional activity substantiated by docking analysis. PLoS One 8, 60567. Kunisue, T., Chen, Z., Louis, G.M.B., Sundaram, R., Hediger, M.L., Sun, L., Kannan, K., 2012. Urinary concentrations of benzophenonetype UV filters in U.S. women and their association with endometriosis. Environ. Sci. Technol. 46, 4624–4632. Kunz, P.Y., Galicia, H.F., Fent, K., 2006. Comparison of in vitro and in vivo estrogenic activity of UV filters in fish. Toxicol. Sci. 90, 349–361. Laurentiis, E.D., Minella, M., Sarakha, M., Marrese, A., Minero, C., Mailhot, G., Brigante, M., Vione, D., 2013. Photochemical processes involving the UV absorber benzophenone-4 (2-hydroxy-4-methoxybenzophenone-5-sulphonic acid) in aqueous solution: reaction pathways and implications for surface waters. Water Res. 47, 5943–5953. Li, Y., Li, M., Shi, J.Q., Yang, X., Wang, Z.Y., 2012. Hepatic antioxidative responses to PCDPSs and estimated short-term biotoxicity in freshwater fish. Aquat. Toxicol. 120–121, 90–98. Li, Z.H., Velisek, J., Zlabek, V., Grabic, R., Machova, J., Kolarova, J., Randak, T., 2010. Hepatic antioxidant status and hematological parameters in rainbow trout, Oncorhynchus mykiss, after chronic exposure to carbamazepine. Chem.-Biol. Interact. 183, 98–104. Liao, C.Y., Kannan, K., 2014. Widespread occurrence of benzophenone-type UV light filters in personal care products from China and the United States: an assessment of human exposure. Environ. Sci. Technol. 48, 4103–4109. Liu, H., Sun, P., Liu, H.X., Yang, S.G., Wang, L.S., Wang, Z.Y., 2015. Acute toxicity of benzophenone-type UV filters for Photobacterium phosphoreum and Daphnia magna: QSAR analysis, interspecies relationship and integrated assessment. Chemosphere 135, 182–188. McCord, J.M., Fridovich, I., 1969. Superoxide dismutase an enzymic function for erythrocuprein (hemocuprein). J. Biol. Chem. 244, 6049–6055. Morohoshi, K., Yamamoto, H., Kamata, R., Shiraishi, F., Koda, T., Morita, M., 2005. Estrogenic activity of 37 components of commercial sunscreen lotions evaluated by in vitro assays. Toxicol. In Vitro 19, 457–469. Mottaleb, M.A., Usenko, S., O’Donnell, J.G., Ramirez, A.J., Brooks, B.W., Chambliss, C. K., 2009. Gas chromatography–mass spectrometry screening methods for select UV filters, synthetic musks, alkylphenols, an antimicrobial agent, and an insect repellent in fish. J. Chromatogr. A 1216, 815–823. Paredes, E., Perez, S., Rodil, R., Quintana, J.B., Beiras, R., 2014. Ecotoxicological evaluation of four UV filters using marine organisms from different trophic levels Isochrysis galbana, Mytilus galloprovincialis, Paracentrotus lividus, and Siriella armata. Chemosphere 104, 44–50.
Puerto, M., Pichardo, S., Jos, A., Prieto, A.I., Sevilla, E., Trias, J.E., Camean, A.M., 2010. Differential oxidative stress responses to pure microcystin-LR and microcystincontaining and non-containing cyanobacterial crude extracts on Caco-2 cells. Toxicon 55, 514–522. Qu, R.J., Feng, M.B., Wang, X.H., Qin, L., Wang, C., Wang, Z.Y., Wang, L.S., 2014. Metal accumulation and oxidative stress biomarkers in liver of freshwater fish Carassius auratus following in vivo exposure to waterborne zinc under different pH values. Aquat. Toxicol. 150, 9–16. Reuter, S., Gupta, S.C., Madan, M., Chaturvedi, M.M., Aggarwal, B.B., 2010. Oxidative stress in flammation and cancer: how are they linked? Free Radical Biol. Med. 49, 1603–1616. Tao, Y.X., Pan, L.Q., Zhang, H., Tian, S.M., 2013. Assessment of the toxicity of organochlorine pesticide endosulfan in clams Ruditapes philippinarum. Ecotoxicol. Environ. Saf. 93, 22–30. Tarazona, I., Chisvert, A., Leon, Z., Salvador, A., 2010. Determination of hydroxylated benzophenone UV filters in sea water samples by dispersive liquid–liquid microextraction followed by gas chromatography–mass spectrometry. J. Chromatogr. A 1217, 4771–4778. Tsai, D.Y., Chen, C.L., Ding, W.H., 2014. Optimization of matrix solid-phase dispersion for the rapid determination of salicylate and benzophenone-type UV absorbing substances in marketed fish. Food Chem. 154, 211–216. Tsui, M.M.P., Leung, H.W., Lam, P.K.S., Murphy, M.B., 2014. Seasonal occurrence, removal efficiencies and preliminary risk assessment of multiple classes of organic UV filters in wastewater treatment plants. Water Res. 53, 58–67. Vela-Soria, F., Ballesteros, O., Zafra-Gomez, A., Ballesteros, L., Navalon, A., 2014. A new method for the determination of benzophenone-UV filters in human serum samples by dispersive liquid–liquid microextraction with liquid chromatography–tandem masss pectrometry. Talanta 121, 97–104. Velma, V., Tchounwou, P.B., 2010. Chromium-induced biochemical, genotoxic and histopathologic effects in liver and kidney of goldfish, Carassius auratus. Mutat. Res.—Genet. Toxicol. Environ. Mutagen. 698, 43–51. Wang, L., Asimakopoulos, A.G., Moon, H.B., Nakata, H., Kannan, K., 2013a. Benzotriazole, benzothiazole, and benzophenone compounds in indoor dust from the United States and East Asian countries. Environ. Sci. Technol. 47, 4752–4759. Wang, L., Kannan, K., 2013b. Characteristic profiles of benzonphenone-3 and its derivatives in urine of children and adults from the United States and China. Environ. Sci. Technol. 47, 12532–12538. Weisbrod, C.J., Kunz, P.Y., Zenker, A.K., Fent, K., 2007. Effects of the UV filter benzophenone-2 on reproduction in fish. Toxicol. Appl. Pharmacol. 225, 255–266. Wu, J.W., Chen, H.C., Ding, W.H., 2013. Ultrasound-assisted dispersive liquid–liquid microextraction plus simultaneous silylation for rapid determination of salicylate and benzophenone-type ultraviolet filters in aqueous samples. J. Chromatogr. A 1302, 20–27. Xiao, M., Wei, D.B., Yin, J.X., Wei, G.H., Du, Y.G., 2013. Transformation mechanism of benzophenone-4 in free chlorine promoted chlorination disinfection. Water Res. 47, 6223–6233. Yang, B., Ying, G.G., 2013. Oxidation of benzophenone-3 during water treatment with ferrate (VI). Water Res. 47, 2458–2466. Zhang, J.F., Shen, H., Wang, X.R., Wu, J.C., Xue, Y.Q., 2004. Effects of chronic exposure of 2,4-dichlorophenol on the antioxidant system in liver of freshwater fish Carassius auratus. Chemosphere 55, 167–174. Zhang, T., Sun, H.W., Qin, X.L., Wu, Q., Zhang, Y.F., Ma, J., Kannan, K., 2013. Benzophenone–type UV filters in urine and blood from children, adults, and pregnant women in China: partitioning between blood and urine as well as maternal and fetal cord blood. Sci. Total Environ. 461–462, 49–55. Zhang, Z.F., Ren, N.Q., Li, Y.F., Kunisue, T., Gao, D.W., Kannan, K., 2011. Determination of benzotriazole and benzophenone UV Filters in sediment and sewage sludge. Environ. Sci. Technol. 45, 3909–3916. Zucchi, S., Bluthgen, N., Ieronimo, A., 2011. The UV-absorder benzophenone-4 alters transcripts of genes involved in hormonal pathways in zebrafish (Danio rerio) eleuthero-embryo s and adult males. Toxicol. Appl. Pharmacol 250, 137–146.
