Ecotoxicology and Environmental Safety 115 (2015) 14–18

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Evaluation of the estrogenic and oxidative stress effects of the UV filter 3-benzophenone in zebrafish (Danio rerio) eleuthero-embryos Gabriela Rodríguez-Fuentes a,n, Juan J. Sandoval-Gío a,b, Anita Arroyo-Silva c, Elsa Noreña-Barroso a, Karla S. Escalante-Herrera d, Francisco Olvera-Espinosa d a

Facultad de Química, Universidad Nacional Autónoma de México, Sisal, Yucatan, Mexico Instituto Tecnológico de Tizimín, Tizimin, Yucatan, Mexico c Benemérita Universidad Autónoma de Puebla, Puebla, Puebla, Mexico d Facultad de Ciencias, Universidad Nacional Autónoma de México, Sisal, Yucatan, Mexico b

art ic l e i nf o

a b s t r a c t

Article history: Received 30 June 2014 Received in revised form 26 January 2015 Accepted 31 January 2015

Personal care products have been detected in superficial waters, representing an environmental risk to the biota. Some studies indicated that 3-benzophenone (3BP) alters hormones, inducing vitellogenesis and having adverse effects on fish reproduction. Other studies have reported generation of free radicals and changes in antioxidant enzymes. Therefore, the aim of the present study was to test acute exposure to 3BP at concentrations within and beyond that found environmentally to provide important toxicological information regarding this chemical. We evaluated the effect of 3BP on vitellogenin 1 (VTG1) gene expression and the transcription of the enzymes catalase (CAT), superoxide dismutase (SOD) or glutathione peroxidase (GPx), which are involved in cellular redox balance. Zebrafish eluthero-embryos (168 hpf) were exposed to 1,10, 100, 1000 mg/L 3BP, in addition to a negative control and a 0.1% ethanol control for 48 h. The results of our study indicated a positive significant correlation between exposure concentrations and VTG1 expression (r ¼0.986, p¼0.0028) but only 1000 mg/L 3BP produced a significant increase from control. Acute exposure showed no significant differences in transcription levels of CAT, SOD or GPx at the tested conditions. Nevertheless, a trend toward increase in GPx expression was observed as a positive significant correlation (r¼ 0.928, p¼ 0.017) was noted. & 2015 Published by Elsevier Inc.

Keywords: Danio rerio 3-Benzophenone Oxidative stress Vitellogenin UV filters

1. Introduction Personal-care products (PCPs) have been detected in superficial waters, representing an environmental risk to the biota (Balmer et al., 2005; Peck, 2006; Mackay and Barnthouse, 2010). Among PCPs, UV filters have a relevant role because they have a diversified spectrum of use because they serve as additives in plastic products, tires, clothes and detergents in addition to their traditional use in cosmetic sunscreen (Schlumpf et al., 2004). Within the wide ranges of UV filters, the most important compounds are benzophenones, cinnamates and crylenes, which have high lipophilicity (Gago-Ferrero et al., 2012). Several authors have documented bioaccumulation of UV filters in aquatic biota. In fish, Nagtegaal et al. (1997) found high concentration of UV filters Abbreviations: (3bp)(3BP), 3-benzophenone; (1BP), 1-bezophenone; (VTG), vitellogenin; (CAT), catalase; (SOD), super oxide dismutase; (GPx), glutathione peroxidase n Correspondence to: Unidad Académica Sisal-UNAM, Av. Colón # 503F X 62 y Reforma Colonia Centro, Mérida, Yucatán 97000, México. E-mail address: [email protected] (G. Rodríguez-Fuentes). http://dx.doi.org/10.1016/j.ecoenv.2015.01.033 0147-6513/& 2015 Published by Elsevier Inc.

in Perca fluviatilis and Rutilus rutilus, similar to levels of other xenobiotics, i.e., DDT. Additionally, bioaccumulation of UV filters has been reported in various aquatic organisms (Buser et al., 2006; Bachelot et al., 2012). It is necessary to emphasize that the ecotoxicological implications of these substances have been insufficiently studied in aquatic organisms. In recent years, the study of their ecological impact has focused on mechanisms of endocrine disruption in fish, e.g., several investigators have recorded estrogenic changes in fish or adverse effects on reproduction or fecundity as result of exposure to UV filters (Schlumpf et al., 2001,, 2004; Giokas et al., 2007; Coronado et al., 2008; DíazCruz and Barceló, 2009). Environmental concentrations of 3-benzophenone (3BP) in natural waters range between 0.8 and 8 mg/L (Balmer et al., 2005). 3BP has been detected in Swiss rivers and lakes at levels of 125 ng/ L and in Japan from 25 to 266 ng/L (Fent et al., 2010; Poiger et al., 2004; Kameda et al., 2011). BP3 has also been found at 700 ng/L and 10,400 ng/L in wastewater treatment plant effluent in Switzerland and US respectively (Balmer et al., 2005). Occurrence of this compound have led to its frequent detection in fish (Balmer et al., 2005; Fent et al., 2010), and in human urine, serum, and

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breast milk in humans (Asimakopoulos et al., 2014; Zhang et al., 2013; Liao and Kannan, 2014). In fish, it has been described that concentrations up to 439 mg/L exhibit multiple hormonal activities at the transcription levels in eleuthero-embryos and adult zebrafish (Blüthgen et al., 2012), Kim et al. (2014) reported that endocrine balance and reproduction performance in Japanese medaka was affected by μg/L concentrations of BP3. Pollutants are able to enhance the production of Reactive Oxygen Species (ROS) and to produce oxidative stress (Lushchak, 2011). Cells use various antioxidant mechanisms to counteract damage caused by ROS and include the enzymes, superoxide dismutase (SOD), catalase (CAT), glutathione peroxidase (GPx) and glutathione reductase (GR). Another system used for protection is the non-enzymatic scavenging of radicals by the reduced form of glutathione and vitamins (C, E, beta-carotene) that directly inactivate ROS (Halliwell, 1999). The involvement of antioxidant enzymes such as SOD, CAT and GPx has been suggested to have an important role in protecting cells from oxidative stress (Xu et al., 2013). Information regarding oxidative stress produced by BP3 is scarce. For example, Hanson et al. (2006) previously reported that 3BP produces ROS in human skin. In addition, Gao et al. (2013) indicated that 3BP treatment (1 mg/L) in Tetrahymena thermophila, a protozoan, increased CAT activity and total glutathione, but they found no changes in GPx or SOD activities. Because 3BP has chemical characteristics that make it suitable for bioaccumulation (Balmer et al., 2005; Rodil et al., 2009), acute exposure to concentrations within and beyond those found environmentally were used in the present study to obtain important toxicological information regarding this chemical. We evaluated the effect of 3BP as a pollutant that may induce vitellogenesis and that may produce changes in the gene expression of CAT, superoxide dismutase SOD and glutathione peroxidase GPx.

2. Materials and methods Adult zebrafish were purchased from a local fish farm. Fish were separated by gender and acclimated for one month in the Unit of Chemistry of the National Autonomous University of Mexico in Sisal, Yucatan, using 40-liter glass aquariums (density of 1 fish/l) with constant aeration. The temperature, pH and conductivity were measured using a multiparameter probe (Hach, USA) 26 72 °C, 8.0 7 0.5 and 800 750 mS/cm, respectively. Fish were maintained in a 14/10 light: darkness photoperiod and were fed two times per day ad libitum with commercial food and once a day with Artemia sp. nauplii. After the acclimation period, fish were transferred into a 40 L tank for reproduction at a 2:1 female/ male ratio, with the water temperature at 28 72 °C (Westerfield, 2000). Spawning was induced by turning on the lights in the morning. Fertilized eggs were collected and washed with clean water. Embryos were cultured for 168 h post-fertilization (hpf) in 5 L aquariums. Eleuthero-embryos were transferred to 10 cm glass Petri dishes with 30 mL of treatment solution; 25 eleuthero-embryos were placed in each dish, and 3 replicates were evaluated per treatment. This study was conducted in accordance with institutional guidelines for the protection of animal welfare. Stock solutions of 3BP (CAS number 131-57-7, Fluka, USA) in ethanol were used to prepare working solutions of 3BP in purified drinking water for the bioassays at nominal 3BP concentrations of 1, 10, 100 and 1000 μg/L. Control (purified drinking water) and solvent control (purified drinking water with 0.1% ethanol) treatments were also tested. At the beginning of the experiment, exposure water samples (50 mL) were collected from every treatment and replicate to determine the real 3BP levels by solid-phase microextraction and gas chromatography–mass spectrometry

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(SPME/GC–MS). A sample aliquot of 10 mL was placed in a 20 mL SPME vial (Supelco, USA) with a screw cap with a PTFE/silicone septum (Supelco, USA). Extraction of 3BP was performed by direct immersion for 30 min at 55 °C with a 65 μm PDMS/DVB SPME fiber (Supelco, USA) and 700–900 rpm magnetic stirring. After the extraction, 3BP was quantified using an Agilent Technologies 6850 Gas Chromatography System equipped with a 5975B mass detector and a Zebron ZB-5MSi capillary column (30 m long, 0.25 mm i.d. and 0.25 μm film thickness, Phenomenex, USA). The inlet temperature was 250 °C, and samples were injected in a splitless mode (1 min purge time) with 10 min of desorption time. The oven temperature program started at 110 °C for 2 min, ramp 1 was 20 °C/min until 170 °C, and ramp 2 was 6 °C/min until 260 °C and held for 5 min; the carrier gas was helium. 3BP was determined in selective ion monitoring (SIM) mode (target ions: 151, 227, 228m/z). The spectra generation frequency was 20 Hz, and the interface and ion source temperatures were 290 and 230 °C, respectively. The MS ionization mode was electron ionization (EI). Calibration solutions were used to quantify the 3BP in the samples. Repeatability tests indicated a variation coefficient of 3.01% and the method detection limit was 0.005 μg/L. After 48 h of exposure, eleuthero-embryos were placed in clean 1.5 mL microcentrifuge tubes with 200 mL of RNA laters (Sigma, USA) and snap frozen in liquid nitrogen. Samples were kept at  80 °C until analysis. Eleuthero-embryos were evaluated for relative gene expression of vitellogenin 1 (VTG1), CAT, SOD and GPx. β-Actin (BAC) was used as the housekeeping gene in all qPCR experiments. No statistical differences in expression were found between treatments for the housekeeping gene used in this experiment. Quantification primers for VTG1, CAT, GPx and BAC were designed from the reported GenBank sequences with accession numbers BC139513.1, NM_130912, BC164790 and BC165823.1, respectively, to obtain amplicons of 100–200 base pairs. SOD quantification primers sequences were taken from Gonzalez et al. (2006). Primer sequences are reported in Table 1. Before qPCR, all amplicons were purified and sequenced to verify primer specificity, and efficiency curves were also run to validate the method. Total RNA was extracted using Gene Elutes Mammalian Total RNA Mike Prep Kit (Sigma, USA). Total RNA concentration was determined by evaluating fluorescence with the Quant-its RNA Assay Kit (Invitrogen, USA). Two hundred nanograms of total RNA was used for the synthesis of cDNA with the iScripts Kit (Biorad, USA); 1 mL of cDNA was used for qPCR with iQs SYBR Green Mix (Biorad, USA). The qPCR was carried out in an IQ5s thermocycler (Biorad, USA) with an initial denaturalization at 95 °C for 3 min, followed by 35 cycles of 30 s at 95 °C, 30 s at 55 °C and 30 s at 72 °C and finally, a melting curve from 95 °C to 50 °C Δ-0.5 °C/20 s. The relative expression of the RNA was calculated using the ΔΔCt method (Livak and Schmittgen, 2001). Non-parametric ANOVA (Kruskal Wallis test) was used to determine differences in gene expression between medians. Dunn's post hoc test was used to determine differences from control values. For all cases, the level of significance was set at p o0.05. Spearman non-parametric correlation was used for data analysis. Statistical analysis was performed using GraphPad Prism 5.0 Table 1 Primer sequences for qPCR for vitellogenin 1 (VTG1), catalase (CAT), superoxide dismutase (SOD) and glutathione peroxidase (GPx) genes in zebrafish larvae exposed to 3-benzophenone. Gene

Forward

Reverse

VTG1 CAT SOD GPx BAC

TGAAGTGCGTCGTATCTTGC CAGGAGCGTTTGGCTACTTC TGAGACACGTCGGAGACC GAAATACGTCCGTCCTGGAA GTGCCATCTACGAGGCTTA

TTCTCAGGGAGAGCAGGTGT ATCTGATGACCCAGCCTCAC TGCCGATCACTCCACAGG TCTCCCATAAGGGACACAGG TCTCAGCTGTGGTGGTGAAG

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Table 2 Nominal and measured concentrations (μg/L ) of 3-benzophenone in water used for zebrafish exposure (n.d.¼ not detected). Nominal concentration Measured concentration

Mean

SD

n.d. n.d. 0.9 8.7 100.7 1045.6

n.d. n.d. 0.1 2.9 3.9 72.9

Replicate 1 Replicate 2 Replicate 3 0 (Control) 0 (Solvent control) 1 10 100 1000

n.d. n.d. 0.9 9.5 101.4 965.8

n.d. n.d. 1.0 5.5 96.5 1062.3

n.d. n.d. 0.8 11.16 104.1 1108.8

(GraphPad Software Inc., USA).

3. Results and discussion The mean (7one standard deviation) measured concentrations of 3BP were 0.9 70.1, 8.7 72.9, 100.7 73.9 and 1045.6 772.9 μg/ L, respectively, for nominal concentrations of 1, 10, 100 and 1000 μg/ L. 3BP was not detected in the control and solvent control water samples. The concentrations of 3BP in the exposure water samples are presented in Table 2. The results indicated that the highest concentration of 3BP (1000 mg/L) exhibited a significant difference in the relative expression of VTG1 with respect to the control (p o0.05) (Fig. 1). A positive significant correlation between exposure concentrations and VTG1 expression (r ¼0.986, p ¼0.0028) was noted. Other authors have indicated that 3BP has estrogenic activity, inducing the synthesis of fish VTG, although the results have not been conclusive. For example, Coronado et al. (2008) found induction of VTG genes in medaka and rainbow trout at 620 and 749 mg/L of 3BP, respectively. Nevertheless, in, zebrafish eleuthero-embryos

(exposed to 8.2–438 mg/L, Blüthgen et al., 2012) and in fathead minnow (exposed to 10–5000 mg/L, Kunz et al., 2006) the induction of VTG gene or the VTG plasma concentration was not significant. Differences between these results may be related to species sensitivity, species developmental stage, exposure concentrations, time and type of exposure. VTG is synthesized in the liver by induction of its coding gene by response elements formed by 17-β estradiol (E2) in the estrogen receptors ER (Clelland and Peng, 2009). ER are the link between E2, gene expression and the chain of mechanisms leading to protein synthesis (Brzozowski et al., 1997). The recognition of the hormone by the receptor is due to the complementarity between the ligand cavity and the nonpolar part of estradiol (Kah, 2009). Unfortunately, this ligand is not very selective and it can relate to several xenoestrogens. Thus, the effect of 3BP is directly related to its interaction with ER. Morohoshi et al. (2005) indicated that benzophenone derivatives have different binding affinities with the estrogen receptor; 1-benzophenone (1BP) has a strong affinity to ER compared to BP3 and other benzophenones. Other authors have also reported that 1BP has stronger estrogenic activity than BP3 in vitro (Kunz et al., 2006; Molina-Molina et al., 2008).This indicates that differences in the biotransformation of this compound may change its toxicity and therefore the metabolism of the organism also needs to be taken into consideration. Blüthgen et al. (2012) reported that in zebrafish eleuthero-embryos, 3BP was not transformed to 1BP, indicating that eleuthero-embryos lack the capability to metabolize 3BP, probably because the metabolizing enzymes are not yet fully active. Kim et al. (2014) measured an increase in 1BP concentrations in water of 3BP exposed Japanese medaka, the presence of 1BP was attributed to 3BP biotransformation. These data indicates that developmental differences in 3BP biotransformation may allow eleuthero-embryos to be more resistant to 3BP than older animals. With respect to the results of antioxidant enzymes

Fig. 1. Transcriptional levels of vitellogenin 1 (VTG1), catalase (CAT), superoxide dismutase (SOD) and glutathione peroxidase (GPx) in zebrafish eleuthero-embryos exposed to 3-benzophenone. Line represents the median, box is the interquartile rages and whiskers depicts minimum and maximum. Significant differences with respect to control (pr 0.05) are indicated with n.

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transcription, there were no significant differences in SOD, CAT or GPx transcription between the eleuthero-embryos of the control group and treatments with 3BP (p 4 0.05) (Fig. 1). Nevertheless, a trend toward increase in GPx expression was observed as a positive significant correlation (r ¼0.928, p ¼0.017) was noted. Increasing numbers of researchers have evaluated changes in antioxidant enzyme activities as biomarkers of oxidative stress in zebrafish (Blahová et al., 2013; Praskova et al., 2014). Others have continued their work with the study of differential expression of genes encoding these enzymes to monitor the impact of chemical pollutants (Woo et al., 2009). Results regarding transcriptional changes of antioxidant enzymes has produced contrasted results, for example, some authors have found that during exposures to different stressors transcription was up-regulated (e.g. Li et al., 2013; Tiedke et al., 2013; Jin et al., 2010; Henrik Hansen et al., 2007). On the contrary, others have found down-regulation of these genes (e.g. Chen et al., 2015; Gonzalez-Rey et al., 2014; Sun et al., 2014). These differences may be attributed to the particular conditions of each experiment, to the pollutant and to the tested species. The limited effect on antioxidant genes found in the present work may be because the organisms do not present oxidative stress at the time samples were taken, indicating that their antioxidant systems are able to deal with ROS production. It is important to consider that there could be also an exposure concentration and time dependency, since it has been previously reported that those variables may have an effect in the expression of antioxidant enzymes genes (Li et al., 2014; Jiang et al., 2014; Woo et al., 2009). In addition, it has to be considered that since it has been previously reported that zebrafish eleuthero embryos lack the ability to biotransform certain pollutants (Blüthgen et al., 2012; Yang et al., 2011) ROS production may be reduced and there is no need to increment antioxidant enzymes.

4. Conclusion The results of our study indicate that 1000 mg/L 3BP produced a significant increase in VTG1 relative expression. The acute exposure to 3BP had no or limited effect in the transcription levels of CAT, SOD and GPx under the tested conditions. Further studies will be conducted to evaluate ROS production, antioxidant enzyme activities/expression and biotransformation of 3BP at different zebrafish developmental stages to fully understand the toxicity of this UV filter.

Acknowledgments This project was funded by the Postdoctoral Fellowship Program of the Mexican National Council of Science and Technology (CONACyT) and by the National Autonomous University of Mexico IACOD Grant # IC200111.

References Asimakopoulos, A.G., Thomaidis, N.S., Kannan, K., 2014. Widespread occurrence of bisphenol A diglycidyl ethers, p-hydroxybenzoic acid esters (parabens), benzophenone type-UV filters, triclosan, and triclocarbanin human urine from Athens, Greece. Sci. Total Environ. 470–471, 1243–1249. Bachelot, M., Li, Z., Munaron, D., Le Gall, P., Casellas, C., Fenet, H., Gomez, E., 2012. Organic UV filter concentrations in marine mussels from french coastal regions. Sci. Total Environ. 420, 273–279. Balmer, M., Buser, H.R., Müller, M.D., Poiger, T., 2005. Occurrence of some organic UV filters in wastewater, surface waters and in fish from Swiss lakes. Environ. Sci. Technol. 39, 953–962. Blahová, J., Plhalová, L., Hostovský, M., Divišová, L., Dobšíková, R., Mikulíková, I.,

17

Štěpánová, S., Svobodová, Z., 2013. Oxidative stress responses in zebrafish Danio rerio after subchronic exposure to atrazine. Food Chem. Toxicol. 61, 82–85. Blüthgen, 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. Brzozowski, A.M., Pike, A.C.W., Dauter, Z., Hubbard, R.E., Bonn, T., Engström, O., Öhman, L., Greene, G.L., Gustafsson, J., Carlquist, M., 1997. Molecular basis of agonism and antagonism in the oestrogen receptor. Nature 389, 753–758. Buser, H., Balmer, M.E., Schmid, P., Kohler, M., 2006. Occurrence of UV filters 4-methylbenzylidene camphor and octocrylene in fish from various swiss rivers with inputs from wastewater treatment plants. Environ. Sci. Technol. 40, 1427–1431. Clelland, E., Peng, C., 2009. Endocrine/paracrine control of zebrafish ovarian development. Mol. Cell. Endocrin. 312, 42–52. Chen, H., Zha, J., Yuan, L., Wang, Z., 2015. Effects of fluoxetine on behavior, antioxidant enzyme systems, and multixenobiotic resistance in the Asian clam Corbicula fluminea. Chemosphere 119, 856–862. 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 (2 hydroxy 4 methoxyphenylmethanone) in fish. Aquat. Toxicol. 90, 182187. DíazCruz, M.S., Barceló, D., 2009. Chemical analysis and ecotoxicological effects of organic UV absorbing compounds in aquatic ecosystems. Trends Anal. Chem. 28, 708717. 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., Díaz-Cruz, M.S., Barceló, D., 2012. An overview of UV-absorbing compounds (organic UV filters) in aquatic biota. Anal. Bioanal. Chem. 404, 2597–2610. Gao, L., Yuan, T., Zhou, C., Cheng, P., Bai, Q., Ao, J., Wang, W., Zhang, H., 2013. Effects of four commonly used UV filters on the growth, cell viability and oxidative stress responses of the Tetrahymena thermophila. Chemosphere 93, 2507–2513. Giokas, D.L., Salvador, A., Chisvert, A., 2007. UV filters: from sunscreens to human body and environment. Trends Anal. Chem. 26, 360374. Gonzalez, P., Baudrimont, M., Boudou, A., Bourdineaud, J.P., 2006. Comparative effects of direct cadmium contamination on gene expression in gills, liver, skeletal muscles and brain of the zebrafish (Danio rerio). Biometals 19, 225–235. Gonzalez-Rey, M., Mattos, J.J., Piazza, C.E., Bainy, A.C.D., Bebianno, M.J., 2014. Effectsof active pharmaceutical ingredients mixtures in mussel Mytilus galloprovincialis. Aquat. Toxicol. 153, 12–26. Halliwell, B., 1999. Antioxidant defense mechanisms: from the beginning to the end (of the beginning). Free Radic. Res. 31, 261–272. Hanson, K.M., Gratton, E., Bardeen, C.J., 2006. Sunscreen enhancement of UV induced reactive oxygen species in the skin. Free Radic. Biol. Med. 41, 12051212. Henrik Hansen, B., Rømma, S., Garmo, ØA., Pedersen, S.A., Olsvik, P.A., Andersen, R. A., 2007. Induction and activity of oxidative stress-related proteins during waterborne Cd/Zn-exposure in brown trout (Salmo trutta). Chemosphere 67, 2241–2249. Jiang, Q., Zhang, W., Tan, H., Pan, D., Yang, Y., Ren, Q., Yang, J., 2014. Analysis of gene expression changes, caused by exposure to nitrite, in metabolic and antioxidant enzymes in the red claw crayfish, Cherax quadricarinatus. Ecotoxicol. Environ. Saf. 104, 423–428. Jin, Y., Zhang, X., Shu, L., Chen, L., Sun, L., Qian, H., Liu, W., Fu, Z., 2010. Oxidative stress response and gene expression with atrazine exposure in adult female zebrafish (Danio rerio). Chemosphere 78, 846–852. Kah, O., 2009. Endocrine targets of the hypothalamus and pituitary In: Bernier, J. Nicholas, Van Der Kraak, Glen, Farrell, Anthony P., Brauner, Colin J. (Eds.), Fish Physiology, vol. 28. Academic Press, Burlington, ISBN: 978-0-12-374631-3, pp. 75–112. Kameda, Y., Kimura, K., Miyazaki, M., 2011. Occurrence and profiles of organic sunblocking agents in surface waters and sediments in Japanese rivers and lakes. Environ. Pollut. 159, 1570–1576. Kim, S., Jung, D., Kho, Y., Choi, K., 2014. Effects of benzophenone-3 exposure on endocrine disruption and reproduction of Japanese medaka (Oryzias latipes) – a two generation exposure study. Aquat. Toxicol. 155, 244–252. Kunz, P.Y., Galicia, H.F., Fent, K., 2006. Comparison of in vitro and in vivo estrogenic activities of UV filters in fish. Toxicol. Sci. 90, 349–361. Li, M., Zheng, Y., Liang, H., Zou, L., Sun, J., Zhang, Y., Qin, F., Liu, S., Wang, Z., 2013. Molecular cloning and characterization of Cat, gpx1 and Cu/Zn-sod genes in Pengze crucian carp (Carassius auratus var . Pengze) and antioxidant enzyme modulation induced by hexavalent chromium in juveniles. Comp. Biochem. Physiol. C 57, 310–321. Li, Z., Chen, L., Wu, Y., Li, P., Li, Y., Ni, Z., 2014. Effects of mercury on oxidative stress and gene expression of potential biomarkers in larvae of the Chinese rare minnow Gobiocypris rarus. Arch. Environ. Contam. Toxicol. 67, 245–251. Liao, C., 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. Livak, K.J., Schmittgen, T.D., 2001. Analysis of relative gene expression data using real-time quantitative PCR and the 2-ΔΔCT method. Methods 25, 402–408. Lushchak, V.I., 2011. Environmentally induced oxidative stress in aquatic animals. Aquat. Toxicol. 101, 13–30. Mackay, D., Barnthouse, L., 2010. Integrated risk assessment of household chemicals and consumer products: addressing concern about triclosan. Integr. Environ. Assess. Manag. 6, 390392. Molina-Molina, J., Escande, A., Pillon, A., Gomez, E., Pakdel, F., Cavaillès, V., Olea, N., Aït-Aïssa, S., Balaguer, P., 2008. Profiling of benzophenone derivatives using fish

18

G. Rodríguez-Fuentes et al. / Ecotoxicology and Environmental Safety 115 (2015) 14–18

and human estrogen receptor-specific in vitro bioassays. Toxicol. Appl. Pharmacol. 232, 384–395. 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. Nagtegaal, M., Ternes, T.A., Bauman, W., Nagel, R., 1997. UV-Filtersubstanzen in wasser and Fischen. Z. Umw. Chem. Ökotoxicol. 9, 79–86. Peck, A.M., 2006. Analytical methods for the determination of persistent ingredients of personal care products in environmental matrices. Anal. Bioanal. Chem. 386, 907939. Poiger, T., Buser, H., Balmer, M.E., Bergqvist, P., Müller, M.D., 2004. Occurrence of UV filter compounds from sunscreens in surface waters: regional mass balance in two Swiss lakes. Chemosphere 55, 951–963. Praskova, E., Plhalova, L., Chromcova, L., Stepanova, S., Bedanova, I., Blahova, J., Hostovsky, M., Skoric, M., Maršálek, P., Voslarova, E., Svobodova, Z., 2014. Effects of subchronic exposure of diclofenac on growth, histopathological changes, and oxidative stress in zebrafish (Danio rerio). Sci. World J. 2014, 5 10.1155/2014/ 645737. Article ID 645737. Rodil, R., Moeder, M., Altenburger, R., Schmitt-Jansen, M., 2009. Photostability and phytotoxicity of selected sunscreen agents and their degradation mixtures in water. Anal. Bioanal. Chem. 395, 1513–1524. Schlumpf, M., Cotton, B., Conscience, M., Haller, V., Steinmann, B., Lichtensteiger, W., 2001. In vitro and in vivo estrogenicity of UV screens. Environ. Health Perspect. 109, 239244. Schlumpf, M., Schmid, P., Durrer, S., Conscience, M., Maerkel, K., Henseler, M.,

Lichtensteiger, W., 2004. Endocrine activity and developmental toxicity of cosmetic UV filters – an update. Toxicology 205, 113–122. Sun, H., Wang, W., Li, J., Yang, Z., 2014. Growth, oxidative stress responses, and gene transcription of juvenile bighead carp (Hypophthalmichthys nobilis) under chronic-term exposure of ammonia. Environ. Toxicol. Chem. 33, 1726–1731. Tiedke, J., Cubuk, C., Burmester, T., 2013. Environmental acidification triggers oxidative stress and enhances globin expression in zebrafish gills. Biochem. Biophys. Res. Commun. 441, 624–629. Westerfield, M., 2000. The Zebrafish Book: A Guide for the Laboratory Use of Zebrafish (Danio rerio). University of Oregon Press, Eugene, OR, USA. Woo, S., Yum, S., Kim, D., Park, H., 2009. Transcripts level responses in a marine medaka (Oryzias javanicus) exposed to organophosphorus pesticide. Comp. Biochem. Physiol. C Toxicol. Phamacol. 149, 427–432. Yang, D., Lauridsen, H., Buels, K., Chi, L.H., La Du, J., Bruun, D.A., Lein, P.J., 2011. Chlorpyrifos-oxon disrupts zebrafish axonal growth and motor behavior. Toxicol. Sci. 121, 146–159. Xu, H., Shao, X., Zhang, Z., Zou, Y., Wu, X., Yang, L., 2013. Oxidative stress and immune related gene expression following exposure to di-n-butyl phthalate and diethyl phthalate in zebrafish embryos. Ecotoxicol. Environ. Saf. 93, 39–44. Zhang, T., Sun, H., Qin, X., Wu, Q., Zhang, Y., Ma, J., Kannan, K., 2013. Benzophenonetype 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.