Science of the Total Environment 505 (2015) 282–289

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Ecotoxicity of ketoprofen, diclofenac, atenolol and their photolysis byproducts in zebrafish (Danio rerio) M.S. Diniz a,⁎, R. Salgado a,b,⁎⁎, V.J. Pereira c,d, G. Carvalho a,c, A. Oehmen a, M.A.M. Reis a, J.P. Noronha a a

REQUIMTE/CQFB, Chemistry Department, FCT, Universidade Nova de Lisboa, 2829-516 Caparica, Portugal ESTS-IPS, Escola Superior de Tecnologia de Setúbal do Instituto Politécnico de Setúbal, Rua Vale de Chaves, Campus do IPS, Estefanilha, 2910-761 Setúbal, Portugal Instituto de Biologia Experimental e Tecnológica (IBET), Av. da República (EAN), 2784-505 Oeiras, Portugal d Instituto de Tecnologia Química e Biológica (ITQB)—Universidade Nova de Lisboa (UNL), Estação Agronómica Nacional, Av. da República, 2780-157 Oeiras, Portugal b c

H I G H L I G H T S • • • • •

Toxicity evaluated for 3 common pharmaceuticals (atenolol, ketoprofen and diclofenac). Toxicity assessed for the pharmaceuticals and UV photolysis by-products in zebrafish. Diclofenac photolysis by-products are more toxic than the parent compound. Ketoprofen and atenolol show stronger oxidative stress response than by-products. UV photolysis should ensure full removal of diclofenac metabolites to avoid toxicity.

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Article history: Received 14 July 2014 Received in revised form 29 September 2014 Accepted 29 September 2014 Available online xxxx Editor: D. Barcelo Keywords: Pharmaceuticals Photolysis By-products Toxicity Zebrafish Oxidative stress

a b s t r a c t The occurrence of pharmaceutical compounds in wastewater treatment plants and surface waters has been detected worldwide, constituting a potential risk for aquatic ecosystems. Adult zebrafish, of both sexes, were exposed to three common pharmaceutical compounds (atenolol, ketoprofen and diclofenac) and their UV photolysis by-products over seven days. The results show that diclofenac was removed to concentrations b LOD after 5 min of UV irradiation. The oxidative stress response of zebrafish to pharmaceuticals and their photolysis by-products was evaluated through oxidative stress enzymes (glutathione-S-transferase, catalase, superoxide dismutase) and lipid peroxidation. Results suggest that the photolysis by-products of diclofenac were more toxic than those from the other compounds tested, showing an increase in GST and CAT levels, which are also supported by higher MDA levels. Overall, the toxicity of waters containing atenolol and ketoprofen was reduced after the parent compounds were transformed by photolysis, whereas the toxicity increased significantly from the by-products generated through diclofenac photolysis. Therefore, diclofenac photolysis would possibly necessitate higher irradiation time to ensure that the associated by-products are completely degraded to harmless form(s). © 2014 Elsevier B.V. All rights reserved.

1. Introduction Pharmaceutically active compounds (PhACs) are a growing, major problem of trace organic contamination in aquatic environments and have been found to be ubiquitously present in wastewater treatment plant (WWTP) effluents across the world (Daughton and Ternes, 1999; Heberer, 2002; Fent et al., 2006; Kolpin et al., 2002). Indeed, ⁎ Corresponding author. Tel.: +351 212 948 571; fax: +351 212 948 550. ⁎⁎ Correspondence to: R. Salgado, REQUIMTE/CQFB, Chemistry Department, FCT, Universidade Nova de Lisboa, 2829-516 Caparica, Portugal. Tel.: +351 212 948 571; fax: +351 212 948 550. E-mail addresses: [email protected] (M.S. Diniz), [email protected] (R. Salgado), [email protected] (V.J. Pereira), [email protected] (G. Carvalho), [email protected] (A. Oehmen), [email protected] (M.A.M. Reis), [email protected] (J.P. Noronha).

http://dx.doi.org/10.1016/j.scitotenv.2014.09.103 0048-9697/© 2014 Elsevier B.V. All rights reserved.

they are increasingly being used in human and veterinary medicine with more than 3000 different substances being used in the European Union (Fent et al., 2006; Christen et al., 2010). These compounds reach WWTP mainly as the result of human excretion and/or disposal of unused medication and are released to the aquatic environments in metabolised and/or un-metabolised forms via WWTP discharges (Heberer, 2002; Cleuvers, 2005; Radjenovic et al., 2009). Municipal WWTPs are a major source of PhACs, showing removal efficiencies ranging from 0 to more than 90% depending on the compound and its persistence (Fent et al., 2006; Kasprzyk-Hordern et al., 2009; Fernández et al., 2014). On-site wastewater treatment systems have been found to have removal efficiencies that are lower than or comparable to conventional WWTPs (Matamoros et al., 2007; Garcia et al., 2013; Du et al., 2014). Overall, WWTPs have been shown to be unable to completely remove PhACs and their by-products.

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Most of the PhACs identified in surface waters are non-steroidal anti-inflammatory drugs (NSAID) commonly used to treat symptoms such as fever, pain conditions and inflammation. Two important compounds in this group include diclofenac and ketoprofen, which are used to suppress inflammation, stronger pain and rheumatic diseases (Fent et al., 2006; Cuklev et al., 2012). Other classes of PhACs that are commonly detected in waters include β-blockers used to treat human hypertension: atenolol is one of the most abundant β-blockers reported (Kasprzyk-Hordern et al., 2009). These PhACs have been frequently detected in waterways as micropollutants (ranging from ng L−1 to μg L−1) that can have potentially harmful environmental effects, even at very low concentrations that are in the ng L− 1 range (Heberer, 2002; Kuemmerer, 2009; Christen et al., 2010; Kuemmerer, 2010; Verlicchi et al., 2012). For instance, ketoprofen and diclofenac in effluents from WWTP are usually found at levels up to 1.0 μg L−1 and several studies have reported toxic effects on fish at these concentrations (Cuklev et al., 2012). The uncertainty surrounding the behaviour of PhACs within WWTPs makes the establishment of strategies to remove PhACs difficult (Fent et al., 2006; Choi et al., 2014). With respect to treatment processes, tertiary treatment approaches have been often proposed to optimise the removal of PhACs from wastewaters (Radjenovic et al., 2009; Jelic et al., 2012; Salgado et al., 2012). Ultraviolet (UV) irradiation is routinely employed for disinfection in WWTPs (Oppenländer, 2003), and has also been demonstrated to reduce the concentration of recalcitrant organics, including PhACs (Canonica et al., 2008; Rosario-Ortiz et al., 2010; Salgado et al., 2012). However, UV irradiation can also generate photolysis by-products that, in some cases, may be more recalcitrant or toxic than the parent compounds (Coelho et al., 2009). Ketoprofen, diclofenac and atenolol have been found to be at least partially degraded during the disinfection process (Salgado et al., 2012). Recently, Salgado et al. (2013) have identified numerous transformation by-products of these three PhACs that were formed during UV photolysis. However, it is of high importance to not only identify such metabolites, but also to establish their relative toxicity with respect to their parent compounds. This is of importance towards designing more effective treatment strategies of PhACs that minimises not only their total discharge to the environment, but also to prevent the discharge of by-products that are more harmful than the parent compounds. Zebrafish (Danio rerio) is a recognised research model for studies in molecular genetics and developmental biology (Belyaeva et al., 2009), and has been referred to as an ideal model for toxicology research (Hill et al., 2005). In addition, there are diverse standard toxicity tests described and recommended by several international organisations (e.g. OECD, U.S.EPA) to test toxicity in adult zebrafish and embryos (Belyaeva et al., 2009; Selderslaghs et al., 2009). Moreover, zebrafish offers several advantages for toxicity testing over other vertebrate species (e.g. easy to house, high fecundity, small size) (Hill et al., 2005). Generally, ecotoxicological assessment of pharmaceuticals has been based on acute toxicity experiments performed by standard tests according to existing guidelines, as outlined above. However, there is very little information about the chronic toxicity, or the bioaccumulation potential of pharmaceuticals in biota and food chains. It should be noted that most of the main questions surrounding this issue are recognized as top research questions by the expert horizon workshop (see Boxall et al., 2012). Biomarkers of oxidative stress can be used to assess the chronic effects caused by several types of environmental contaminants on organisms. Oxidative stress is usually defined as an imbalance between reactive oxygen species (ROS) production and antioxidant defences in living organisms (Livingstone, 2001; Oruc and Uner, 2002). The ROS are induced in cells by a diversity of chemical compounds (e.g. metals, pesticides, pharmaceuticals, PAHs) during cells' metabolism and can cause oxidation of proteins and lipids, alterations in gene expression, and alterations in the status of cell redox. ROS are able to cause damage to organisms' cells and have been associated with several pathologies

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(Gong et al., 1997; Livingstone, 2001, 2003). Nonetheless, organisms have developed complex antioxidant systems to respond and minimise the deleterious effects of ROS in cells before they damage cellular components, by catalysing the decomposition of oxidants and free radicals (Gong et al., 1997), thus protecting the functions of organisms from oxidative stress. Mechanisms of antioxidant defence in fish involve their enzyme system and low molecular weight antioxidants, similar to those in mammals, although the specific isoforms of enzymes in various fish species have not been well identified (Di Giulio and Meyer, 2008). Superoxide dismutase (SOD), catalase (CAT), glutathione peroxidase (GPx), and glutathione-s-transferase (GST) are the main antioxidant enzymes and important indicators of oxidative stress. Reduced glutathione (GSH) plays a key role in non-enzymatic antioxidant defence (Sevcikova et al., 2011). Furthermore, the by-products of oxidative stress can be assessed through the malondialdehyde (MDA) assay, which measures lipid peroxidation and is considered to be an indication of oxidative damage to cell membranes (Buege and Aust, 1978). In this study, we assessed the toxicity of ketoprofen, diclofenac and atenolol and their photodegradation by-products in exposed zebrafish through evaluating oxidative stress markers (GST, SOD, CAT, MDA). This enabled a direct comparison between the relative ecotoxicity of the parent compound vs metabolite, in order to determine the potential for UV photolysis to serve as an effective treatment technology to remove these commonly observed compounds from WWTP, without generating harmful by-products. 2. Materials and methods 2.1. Reagents The PhACs used in this study were atenolol, diclofenac and ketoprofen, purchased from Sigma-Aldrich, Portugal. For each compound, a standard solution of 1.0 mg mL−1 in methanol was prepared and stored at 4 °C. The mobile phases used in high performance liquid chromatography (HPLC) were acetonitrile (HPLC grade, Panreac, Portugal) and ultra-pure water obtained from a Milli-Q50 water purification system (Millipore, Bedford, USA), acidified with formic acid (analytical grade, Merck, Portugal). 2.2. Photolysis with a medium pressure (MP) UV lamp The photodegradation tests were carried out in a pear-shaped glass reactor with a volume of 300 mL, using a medium pressure (MP) Hg lamp, Heraeus Noblelight model TQ 150 (nominal power 150 W), which emits radiation between 200 and 450 nm. Low pressure and medium pressure lamps are widely used to achieve disinfection and degradation of xenobiotics. Even though low pressure lamps have a higher germicidal efficiency, a medium pressure (MP) mercury lamp was chosen in this study due to its ability to emit light over a wider range of wavelengths that will lead to a higher overlay between the emission spectra of the lamp and the light absorption of the pharmaceutical compounds and a consequent higher degree of degradation expected. These lamps have also been reported as an effective alternative to low pressure (LP) lamps in compact treatment systems, since the UV intensity per lamp is higher than in LP systems (Oppenlander 2003). The lamp was covered with a quartz cooling jacket, where pure water (with negligible light absorption in the wavelength range of emitted radiation) was used as an optical filter and to maintain a temperature of 25 ± 1 °C. The UV photon fluence rate determined for ketoprofen, atenolol and diclofenac was 0.000085, 0.000579 and 0.002444 einstein m−2 s−1, respectively (Salgado et al., 2013). Three hundred millilitres of tap water, previously aerated for 5 h to remove residual chlorine, was spiked with atenolol, ketoprofen or diclofenac, separately, to get a concentration of 1 mg L− 1 of each

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compound. During the irradiation, the compounds were mixed by bubbling air in the UV reactor. The MP UV irradiation was carried out under the same conditions as employed by Salgado et al. (2013), and all tests were performed in duplicate. Irradiated samples were taken for the toxicity evaluation at the times that corresponded to the production of the highest area of the identified by-products (Salgado et al., 2013): (i) after 7.5 min and 60 min UV exposure for ketoprofen; (ii) after 1.5 min and 5 min UV exposure for diclofenac; and (iii) after 17.5 min and 90 min UV exposure for atenolol. Two millilitre samples were also taken at these respective times during the UV photolysis experiments performed in this study, in order to assess the relative area of chromatographic peaks of the PhACs through HPLC analysis. A blank tap water sample (previously aerated for 5 h), without any PhAC spike was also irradiated for comparison purposes, where samples were collected at each of the aforementioned time intervals. 2.3. HPLC-DAD-MS analysis HPLC analysis with a diode array detector (DAD) was used to monitor diclofenac, ketoprofen, atenolol and metabolite concentrations generated in the UV irradiation experiments. HPLC-DAD was carried out in a HPLC system (Waters) coupled with a pump and controller (Waters 600), an in-line degasser (X-Act-4 channels, Jour Research), an auto-sampler (Waters 717 plus) and a photodiode array detector (Waters 996). Reverse-phase chromatography (LiChroCART 250-4 Purospher Star RP18 endcapped, 5 μm, column, Merck) of the samples (injection volume of 50 μL) was performed using a degassed mobile phase with 70% water with 0.01% formic acid and 30% acetonitrile with a 0.6 mL min− 1 flow rate, with DAD detection from 200 to 400 nm. The chromatograms were acquired with a MassLinxTM software data acquisition system. The (ESI+)MS was carried out with a quadropole VG Platform (Micromass, UK) spectrometer equipped with an electrospray ionisation (ESI) source operating in positive mode. An accurate splitter (split ratio of 1:10) was used between the HPLC column and the mass spectrometer. Capillary temperature was kept between 100 °C and 120 °C, using a scanning cone voltage from 35 to 100 V and capillary voltage of 3.5 kV. Nitrogen was used as drying and nebulising gas at 300 mL min−1 and 10 mL min−1, respectively. The mass/charge spectrum range used was 100–450 amu with a MassLynxTM software data acquisition system. Limits of detection of ketoprofen, diclofenac and atenolol were 23, 129 and 78 μg L−1, respectively. 2.4. Exposure assays D. rerio were obtained from a commercial supplier (Mil Aquários, Portugal) and acclimated to laboratory conditions throughout two weeks before assays. The fish were housed in a 100 L volume glass aquarium; a closed circuit system with filtered de-chlorinated tap water, at pH 7.2 ± 0.1, temperature of 20 ± 1 °C, a photoperiod of 12 h light and 12 h dark, as well as continuous aeration (N6 mg O2 L−1). Afterwards, adult fish (n = 72; 0.3 ± 0.1 g; 2.1 ± 0.4 cm standard length), of both sexes, were randomly distributed (since it was not possible to distinguish sex by external observation) into 9 polystyrene test tanks (0.3 L water volume). Three additional tanks with clean tap water subjected to 0, 5 and 90 min of UV irradiation were used as controls (control, control t5 and control t90). Fish were exposed to 1.0 mg L−1 of the different target pharmaceutical compounds (atenolol, ketoprofen and diclofenac) and their photolysis by-products (Salgado et al., 2013), as detailed in Table S1 (supplementary material). The concentration of 1.0 mg L−1 was chosen since it allowed the detection of the metabolites generated from UV radiation, whilst not approaching the mortality threshold for D. rerio, which has been found to be over 2 orders of magnitude higher (Praskova et al., 2011a,b). The experiment was conducted in duplicate for seven days, and fish were fed daily with commercial flakes of dry food (Tetra brand). Tanks were monitored for pH, temperature, and

ammonia, and cumulative mortality was registered. At the end of the experimental period, the fish were sampled, decapitated and dissected to remove the whole visceral mass, which was immediately stored at −80 °C until analysis. 2.5. Biochemical assays 2.5.1. Sample treatment Samples from the whole visceral mass were homogenised on-ice in 2 mL of cold buffer (100 mM potassium phosphate from Sigma-Aldrich, Germany; containing 2 mM of EDTA from Riedel-Haën, Germany) at pH 7.0. Tissue homogenates were centrifuged at 10,000 ×g for 15 min at 4 °C and stored at −80 °C for further analysis. The results were normalised to the total protein mass (mg) determined by the Bradford method (Bradford, 1976). 2.5.2. Glutathione-S-transferase (GST) activity Total GST activity (EC 2.5.1.18) was determined according to the procedure described by Habig et al. (1974) by measuring the formation of the conjugate of glutathione (GSH) and l-chloro-2,4-dinitrobenzene (CDNB). Briefly, 180 μL of substrate solution (Dulbecco's Phosphate Buffered Saline with 200 mM reduced L-glutathione and 100 mM CDNB all from Sigma-Aldrich, Germany) was added to 20 μL of GST standard or sample into each well of a 96-well microplate (Nunc-Roskilde, Denmark). The total enzyme activity was determined at 340 nm by recording the absorbance at every minute for 6 min, using a microplate reader (BioRad Benchmark, USA). Equine liver GST (Sigma-Aldrich, Germany) was used as standard and positive control. The change in absorbance per minute (ΔA340) was estimated and the reaction rate at 340 nm was determined using a CDNB extinction coefficient of 0.0096 μM−1 cm−1. The results are expressed in relation to total protein concentration of the sample (nmol min−1 mg−1 total protein). 2.5.3. Superoxide dismutase (SOD) The SOD (EC 1.15.1.1) determination followed the nitroblue tetrazolium (NBT) method, adapted from Sun et al. (1988). In this method, superoxide radicals (∙O− 2 ) are generated by the reaction of xanthine with xanthine-oxidase (XOD), and reduce NBT to formazan, which can be assessed spectrophotometrically at 560 nm. SOD competes with NBT for the dismutation of ∙ O− 2 , inhibiting its reduction. The inhibition level is used as a measure of SOD activity. SOD from bovine erythrocytes (Sigma-Aldrich, Germany) was used as standard and positive control. The assay was performed using a 96-well microplate (Nunc-Roskilde, Denmark), adding to each well 200 μL of 50 mM phosphate buffer (pH 8.0) (Sigma-Aldrich, Germany), 10 μL of 3 mM EDTA (RiedelHaën, Germany), 10 μL of 3 mM xanthine (Sigma-Aldrich, Germany), 10 μL of 0.75 mM NBT (Sigma-Aldrich, Germany) and 10 μL of SOD standard or sample. The reaction was started with the addition of 100 mU XOD (Sigma-Aldrich, Germany) and the absorbance at 560 nm was recorded every minute for 5 min, using a plate reader (BioRad Benchmark, USA). Negative controls included all components except SOD or sample, producing a maximal increase in absorbance at 560 nm, which allowed determining the inhibition percentage per minute, caused by SOD activity. The total SOD activity is expressed in units/ mg of protein, where one unit is equivalent to the SOD activity that causes 50% inhibition of the reaction rate without SOD. The SOD results are expressed as the % inhibition. 2.5.4. Catalase (CAT) The CAT activity (EC 1.11.1.6) was measured according to a spectrophotometric method adapted from Aebi (1984), following the decrease in absorbance at 240 nm by H2O2 consumption. A substrate solution of 0.036% (w/w) hydrogen peroxide was prepared in 0.05 M potassium phosphate buffer (Sigma-Aldrich, Germany) at pH 7.0 and 25 °C, using hydrogen peroxide (30% (w/w) from Sigma-Aldrich, Germany). For method validation and positive control, standard bovine liver catalase

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(Sigma-Aldrich, Germany) was used. Briefly, 0.1 mL of standard bovine liver catalase or sample (from the previously homogenised tissues) was added to 2.9 mL of substrate solution into a quartz cuvette and the absorbance (240 nm) was measured every 15 s for 2 min (at 25 °C, pH 7.0 and path length 10 mm), using a spectrophotometer (Unicam Helios, Portugal). The change in absorbance per minute (ΔA240) was estimated and the reaction rate was determined using a H2O2 extinction coefficient of 40 M−1 cm− 1. The results are expressed in relation to the total protein mass of the sample (nmol min−1 mg−1 total protein). 2.5.5. Lipid peroxide assay (MDA content) The lipid peroxide assay was adapted from the thiobarbituric acid reactive substances (TBARS) protocol (Uchiyama and Mihara, 1978). Briefly, 5 μL of each sample was added to 45 μL of 50 mM monobasic sodium phosphate buffer. Then, 12.5 μL of SDS 8.1%, 93.5 μL of trichloroacetic acid (20%, pH = 3.5) and 93.5 μL of thiobarbituric acid (1%) were added to each microtube. To this mixture, 50.5 μL of MQ-grade ultrapure water were added and microtubes were agitated in a vortex for 30 s. The lids were punctured with a needle and microtubes were incubated in boiling water for 10 min and subsequently placed on ice for a few minutes to cool. Then, 62.5 μL of MQ-grade ultrapure water and 312.5 μL of n-butanol pyridine (15:1, v/v) were added. Microtubes were then centrifuged at 5000 ×g for 5 min. Duplicates of 150 μL of the supernatant of each reaction were added into each well of a 96-well microplate and absorbance was read at 530 nm. To quantify the lipid peroxides, an eight-point calibration curve (0–0.3 μM TBARS) was constructed using malondialdehyde bis(dimethylacetal) (MDA) (from Merck) standards. 2.6. Statistical analysis Statistical analysis was performed with Statistica software (Statistica version 8.0; Statsoft Inc., Tulsa, OK, USA, 2007) at a significance level of 5%. After failure to demonstrate statistical assumptions such as homogeneity of variances by Levene's test, the nonparametric Mann–Whitney U test was used to determine differences in PhACs and their metabolites before and after UV photolysis and between the treated and the control fish samples. The non-parametric Spearman R was used to analyse the correlation amongst antioxidant enzymes and lipid peroxidation. 3. Results 3.1. Fate of PhAC and photolysis by-products during ecotoxicological tests The abundance of ketoprofen, diclofenac and atenolol and their respective metabolites before and after the ecotoxicity tests with D. rerio is shown in Fig. 1a–c, respectively. From Fig. 1, it can be observed that most compounds were found in a similar abundance at the beginning and end of the toxicity tests (within a given UV exposure time), showing that the PhACs did not tend to accumulate in D. rerio through e.g. adsorption/absorption or be removed by other mechanisms. The primary exception was atenolol (Fig. 1c), where 61% to 92% removal was observed during the tests with D. rerio. The previously identified metabolites by Salgado et al. (2013) are shown in Table S1 of the supplementary material. As shown in Fig. 1, a higher number of metabolites were usually observed in the intermediate sampling point of UV photolysis, particularly for ketoprofen and diclofenac. 3.2. Effects of the UV irradiated PhAC in D. rerio 3.2.1. Mortality No significant mortality was observed during the experimental period for all tested PhAC (one fish died in the controls and another one that was exposed to atenolol).

Fig. 1. Peak area (mean ± s.d.) of ketoprofen (a), diclofenac (b) atenolol (c) and their by-products generated through UV photolysis, measured at the beginning and end of the toxicity tests. Asterisks mark significant differences in PhACs and their metabolites before and after UV photolysis (p b 0.05).

3.2.2. Glutathione-S-transferase (GST) activity Average concentrations of the cytosolic GST total activity measured in the whole visceral mass from D. rerio are presented in Fig. 2. The highest concentrations of GST activity (19.44 nmol min−1 mg−1 total protein) were measured in fish exposed to 1.0 mg L−1 diclofenac that was irradiated for 1.5 min, whereas the lowest concentrations (4.01 nmol min−1 mg−1 total protein) were found in controls. Regarding GST activity, a significant decreasing trend (p b 0.05) was observed

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Fig. 2. GST (mean ± s.d.) in D. rerio exposed to PhACs irradiated for different times. Significant differences if letters are different (p b 0.05).

for atenolol after irradiation for 90 min and for ketoprofen following 7.5 min irradiation, whereas for diclofenac a significant increase (p b 0.05) was found after irradiation for 1.5 and 5 min. Moreover, a significant but low positive correlation (r = 0.42; p b 0.05) was found between GST activities and lipid peroxidation (see 3.2.5).

3.2.3. Superoxide dismutase (SOD) Average percentages of the cytosolic SOD (% inhibition) determined in the whole visceral mass from D. rerio are presented in Fig. 3. The lowest percentage of SOD inhibition (16%) was determined in fish exposed to 1.0 mg L−1 atenolol that was exposed to UV irradiation for 90 min, whereas the highest percentage of inhibition (63%) was found in fish exposed to 1.0 mg L− 1 ketoprofen without UV irradiation. A significant decreasing trend (p b 0.05) in the % inhibition was observed

for atenolol and ketoprofen following irradiation. No significant correlations were found between SOD and lipid peroxidation (see 3.2.5).

3.2.4. Catalase (CAT) Average concentrations of the catalase activity measured in the whole visceral mass from D. rerio are presented in Fig. 4. The highest concentration of CAT activity (98.17 nmol min−1 mg−1 total protein; Fig. 4) was measured in fish exposed to 1.0 mg L−1 diclofenac after 1.5 minute exposure to UV, whereas the lowest concentrations (12.40 nmol min−1 mg−1 total protein) were determined in controls. Concerning catalase activity, a significant decreasing trend (p b 0.05) was observed after atenolol or ketoprofen were irradiated, whilst a significant increase (p b 0.05) was observed for diclofenac after irradiation. In addition, significant strong correlations (r = 0.68; p b 0.05) were

Fig. 3. Percentage inhibition of SOD (mean ± s.d.) in D. rerio exposed to PhACs irradiated for different times. Significant differences if letters are different (p b 0.05).

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Fig. 4. Catalase (mean ± s.d.) in D. rerio exposed to PhACs irradiated for different times. Significant differences if letters are different (p b 0.05).

found between CAT and lipid peroxidation and between CAT and GST (0.69; p b 0.05) (see 3.2.5).

3.2.5. Lipid peroxidation (MDA concentration) The average MDA concentrations measured in the whole visceral mass from D. rerio are presented in Fig. 5. The highest concentration of MDA (20.83 nmol mg−1 total protein) was measured in fish exposed to diclofenac after 1.5 min of UV irradiation, whereas the lowest concentrations (2.92 nmol mg−1 total protein) were determined in controls. Concerning MDA levels, a significant decreasing trend (p b 0.05) was observed in irradiated atenolol and ketoprofen whilst a significant increase (p b 0.05) was observed for diclofenac after irradiation.

4. Discussion The HPLC-DAD-MS analysis confirmed the presence of parent compounds (ketoprofen, atenolol and diclofenac) in water samples exposed to UV photolysis, as well as their metabolites, which increased with exposure time to UV irradiation (see Fig. 1). As expected, a higher number of metabolites were usually observed in the intermediate sampling point of UV photolysis, particularly for ketoprofen and diclofenac, which is also consistent with previous observations carried out by Salgado et al. (2013) and from Vogna et al. (2004). Despite the use of various removal methods in WWTP, many pharmaceutical compounds and by-products have been identified in treated wastewaters, ground and surface waters worldwide (Heberer, 2002; Behera

Fig. 5. MDA concentration (mean ± s.d.) in D. rerio exposed to different pharmaceutical compounds. Significant differences if letters are different (p b 0.05).

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et al., 2011; Fang et al., 2012; Cardoso et al., 2014; Moreno-González et al., 2014). Throughout the exposure period no mortality was recorded, suggesting that the tested pharmaceuticals do not compromise the survival of the organisms. This is in agreement with literature studies showing that compounds such as diclofenac, ketoprofen and atenolol do not exhibit acute toxicity or lethal effects on organisms at the concentration levels (1.0 mg L−1) tested (Parolini et al., 2011; Prášková et al., 2013). Oxidative stress may produce DNA damage, enzymatic inactivation, and peroxidation of cell constituents, especially lipid peroxidation when antioxidant defences are impaired or overcome (Himmelfarb and Hakim, 2003). In the present work, oxidative stress response was observed indicating that the pharmaceuticals (especially diclofenac) and its photolysis by-products were, to some extent, able to cause low to moderate toxicity, as shown by increasing activities of oxidative stress enzymes and lipid peroxidation. The GST has a dual protective role associated to phase II detoxification enzymes by catalysing the conjugation of the reduced form of glutathione (GSH) to electrophilic compounds, but it has also a protective role by using GSH in the reduction of a wide range of organic hydroperoxides (van der Oost et al., 2003). The results from GST activities in fish exposed to diclofenac and its photolysis by-products show a significant increase in comparison to controls. In effect, diclofenac has been identified as an important pharmaceutically active compound present in the aquatic environment (Letzel et al., 2009) and classified as harmful to aquatic biota (Carlsson et al., 2006). According to Brandao et al. (2013) an increase in GST levels suggests an oxidative stress-related adaptive response. Additionally, a significant increase in GST activity for diclofenac metabolites in comparison to the parent compound was also observed, suggesting that they may present more toxicity than the parent compound. A Microtox assay (Vibrio fischeri) from Calza et al. (2006) showed that the photocatalytic degradation of diclofenac via TiO2-produced by-products was capable of causing more toxicity than the parent compound. However, the toxicity decreased following degradation of the intermediates. Nonetheless, the by-products observed in their study were different to those observed in the present work (see Table S1). Catalase results for fish exposed to diclofenac photodegradation products showed a significant increase, which is in agreement with the increase observed in LPO (MDA content). In effect, the LPO results are compatible with oxidative stress, affecting cell membranes, which was more pronounced in organisms exposed to the photolysis byproducts of diclofenac. However, no changes were recorded for SOD in fish exposed to irradiated diclofenac. Other studies are also in accordance with the potential toxicity of diclofenac, where van den Brandhof and Montforts (2010) reported effects on growth retardation in zebrafish exposed to diclofenac (N1.5 mg L−1), Hallare et al. (2004) showed a delay in the hatching time of exposed zebrafish (1 and 2 mg diclofenac L−1) and Ferrari et al. (2003) observed a 10 day NOEC (at 4 mg L−1) for the same species. Schwaiger et al. (2004) showed that long term effects resulted in a lowest observed effect concentration (LOEC), of 1 to 5 μg diclofenac L−1 in rainbow trout. Moreover, studies from Hoeger et al. (2005) demonstrated that diclofenac residues have the potential to adversely affect various tissues (gills and kidneys) in brown trout at concentrations close to those regularly found in surface waters. In addition, Gonzalez-Rey and Bebianno (2014) exposed mussels (Mytilus galloprovincialis) to environmentally relevant concentrations of diclofenac (250 ng L−1) over 15 days, showing induction of SOD in gills and increased CAT activities and LPO in mussel's digestive glands, but GST levels remained unchanged. Thus, the different studies suggest that diclofenac toxicity effects may differ with the species, exposure time and life cycle stage. In the present study, the results from GST activities revealed a slight decrease in fish exposed to atenolol and ketoprofen photolysis by-products as compared to the parent compounds, suggesting reduced effects on the organisms. Catalase and SOD (% inhibition) results showed

a significant increase for fish exposed to atenolol and ketoprofen in comparison to controls, followed by a decrease with respect to the exposure of their photolysis by-products in comparison to the parent compounds. CAT activities suggested that the formation of H2O2 was not favoured by the products of phototransformation (except for diclofenac), but the rise observed for parent compounds indicated the production of H2O2. These findings suggest that ketoprofen and atenolol and their by-products present lower toxicity than diclofenac and its by-products, which is in agreement with other studies. For example, Sanderson et al. (2003) reported an EC50 in fish for ketoprofen of 32 mg L−1. Winter et al. (2008) exposed fathead minnows (Pimephales promelas) to atenolol (0.1–10 mg L−1) and no significant effects on viability, hatching, or growth were observed. It is known that oxidative stress is attenuated by an antioxidant defence system such as CAT or SOD enzymes that play a crucial role in maintaining cell homeostasis, by maintaining a relatively low rate of the reactive and detrimental hydroxyl radicals (Celik et al., 2006). The significant positive correlations found between enzyme activities and LPO reinforce both the protective role of these antioxidant enzymes, but also the adverse effects of tested PhACs to cell membranes. Despite the significant oxidative stress responses obtained, mainly for diclofenac and its by-products, tested concentrations are higher than typically observed in the environment. However, the present study allows us to highlight the importance of studying the effects of PhACs on aquatic biota and also the relevance of perceiving the effects of their by-products on fish. In addition, the effects of complex mixtures that occur in the real environment and the long-term exposure effects both to PhACs and their transformation products are largely unknown and may cause harmful effects. These aspects require further study. These results could impact future operational strategies for maximising PhAC removal in WWTPs. Considering that the byproducts of diclofenac phototransformation seemed to cause a significantly higher oxidative stress response than the parent compound, employing UV photolysis would likely require increased exposure time as compared to that employed in this study, to potentially promote further degradation of the parent compound and photolysis byproducts. Salgado et al. (2012) have previously observed that both atenolol and ketoprofen are highly biodegradable, whilst diclofenac displays low biodegradability in WWTPs. Since diclofenac is often one of the most frequent and most abundant PhACs in WWTP effluents, sufficient irradiation time should be applied to achieve the degradation of the by-products formed from this compound. 5. Conclusions This study investigated the ecotoxicity of three common pharmaceuticals (atenolol, ketoprofen and diclofenac) and their UV photolysis by-products on zebrafish. The results from four separate tests assessing the impact of the compounds on oxidative stress in the zebrafish showed that ketoprofen and atenolol presented a stronger oxidative stress response than their photolysis by-products, whilst those from diclofenac were significantly more toxic than the parent compound. The application of UV photolysis for diclofenac removal should therefore ensure the degradation of the by-products formed, in order to prevent an increase in the toxicity of the effluent. Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2014.09.103. Acknowledgements The authors acknowledge FCT/MCTES through the project Pest-C/ EQB/LA0006/2013 and project PTDC/AAC-AMB/113091/2009. References Aebi H. Catalase invitro. Methods Enzymol 1984;105:121–6.

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