Response of gene expression in zebrafish exposed to pharmaceutical mixtures: Implications for environmental risk

Ecotoxicology and Environmental Safety 142 (2017) 471–479

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Response of gene expression in zebrafish exposed to pharmaceutical mixtures: Implications for environmental risk Gabriela V. Aguirre-Martíneza,b,c,h, Theodore B. Henrya,e,f,g,h

⁎,1

MARK

, Helena C. Reinardya,e,h, M. Laura Martín-Díazc,d,h,

a

School of Biomedical and Biological Science, 411 Davy Building, University of Plymouth, Drake Circus, Plymouth PL4 8AA, United Kingdom Health Science Faculty, Arturo Prat University, Casilla 121, 1110939 Iquique, Chile c Andalusian Center of Marine Science and Technology (CACYTMAR), Campus Universitario Puerto Real, 11510 Puerto Real, Cádiz, Spain d Facultad Ciencias del Mar y Ambientales, Universidad de Cádiz, Campus de Excelencia Internacional del Mar (CEIMAR), Polígono Río San Pedro s/n. P. Real, Cádiz, Spain e Department of Arctic Technology, UNIS, Longyearbyen N-9171, Norway f School of Life Sciences, Heriot-Watt University, 3.05 William Perkin Building, Edinburgh EH14 4AS, United Kingdom g Center for Environmental Biotechnology, University of Tennessee, Knoxville TN 37996, USA h Department of Forestry, Wildlife and Fisheries, University of Tennessee, Knoxville, TN 37996, USA b

A R T I C L E I N F O

A B S T R A C T

Keywords: Gene expression Toxicogenomics Mixtures Q-PCR Fish Pharmaceuticals

Complex mixtures of pharmaceutical chemicals in surface waters indicate potential for mixture effects in aquatic organisms. The objective of the present study was to evaluate whether effects on target gene expression and enzymatic activity of individual substances at environmentally relevant concentrations were additive when mixed. Expression of zebrafish cytochrome P4501A (cyp1a) and vitellogenin (vtg) genes as well as activity of ethoxyresorufin-O-deethylase (EROD) were analyzed after exposure (96 h) to caffeine-Caf, ibuprofen-Ibu, and carbamazepine-Cbz (0.05 and 5 µM), tamoxifen-Tmx (0.003 and 0.3 µM), and after exposure to pharmaceutical mixtures (low mix: 0.05 µM of Caf, Ibu, Cbz and 0.003 µM of Tmx, and high mix: 5 µM of Caf, Ibu, Cbz and 0.3 µM of Tmx). Pharmaceuticals tested individually caused significant down regulation of both cyp1a and vtg, but EROD activity was not affected. Exposure to low mix did not cause a significant change in gene expression; however, the high mix caused significant up-regulation of cyp1a but did not affect vtg expression. Up-regulation of cyp1a was consistent with induction of EROD activity in larvae exposed to high mix. The complex mixture induced different responses than those observed by the individual substances. Additive toxicity was not supported, and results indicate the need to evaluate complex mixtures rather than models based on individual effects, since in environment drugs are not found in isolation and the effects of their mixtures is poorly understood.

1. Introduction Pharmaceutical substances are constantly released to the aquatic environment principally through municipal effluents and wastewater treatment plants leading to chronic exposure in aquatic organisms (Brain et al., 2004; Halling-Sørensen et al., 1998; Jones et al., 2005; Nakada et al., 2006). Discharges of these substances are likely to increase in the future because of increases in their production, human population growth, and demographic shifts towards higher proportions of older people that use greater amounts of pharmaceuticals (Daughton and Ternes, 1999). Although present at low concentrations at the ng L−1 to µg L−1 range (Gómez et al., 2007; Gros et al., 2007; Ternes,

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1998; Thomas and Foster, 2004), pharmaceutical substances are found simultaneously as complex mixtures that have unknown and difficult to evaluate effects on aquatic biota (Blasco and DelValls, 2008; Gagné et al., 2006; Lara-Martín et al., 2014; Yoon et al., 2010). It is possible that individual substances can act in a synergistic or additive manner, which suggests management for environmental protection should take into consideration mixtures of substances rather than models based on individual effects. Pharmaceuticals are an example of substances for which additive effects on toxicity in aquatic organisms have been observed (Cleuvers et al., 2003; Christensen et al., 2007; Henry and Black, 2007). Pharmaceuticals may induce cytochrome P450 enzyme activity (CYP1A) by binding to the aryl hydro-

Corresponding author. E-mail addresses: [email protected] (G.V. Aguirre-Martínez), [email protected] (T.B. Henry). Present Address: Faculty of Health Science. Arturo Prat University, casilla 121,1110939 Iquique, Chile.

http://dx.doi.org/10.1016/j.ecoenv.2017.04.038 Received 14 November 2016; Received in revised form 15 April 2017; Accepted 17 April 2017 0147-6513/ © 2017 Elsevier Inc. All rights reserved.

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Table 1 Features of the pharmaceuticals selected in this study. Molecular weight (g mol−1)

Log Kow

Group: Mode of Action

Caffeine (58–08–2)

194.19

−0.07a

Stimulant: Inhibitory neurotransmitter that suppresses activity in the central nervous system.

Ibuprofen (15687–27–1)

206.28

3.97a

Anti-inflammatory: Inhibit the enzyme cyclooxygenase.

Carbamazepine (298–46–4)

236.27

2.45a

Anticonvulsant: Stabilizes the inactivated state of voltage-gated sodium channels.

Tamoxifen (10540–29–1)

371.51

6.3b

Anticarcinogenic: anti-estrogen (inhibiting agent) in mammary tissue. Estrogen (stimulating agent) in cholesterol metabolism, bone density, and cell proliferation

Pharmaceutical (Cas Number)

Structure

substances in a mixture are below levels that cause estrogenic effects as single substances, their combined effect within a mixture can be sufficient to induce estrogenicity in fish (Brian et al., 2005; Ketan and Collins, 2007). Consequently, analysis of the potential for substances to influence fish vitellogenin (vtg or VG) is frequently used as a biomarker of estrogenic activity (Filby et al., 2007). Although numerous studies have examined toxicological impacts of pharmaceuticals, the biological activity of these substances in non-target organisms remain uncertain, and interactions among pharmaceuticals when present in mixtures can be considerably more complicated than that of additive toxicity based on a single mechanism of action. The toxicity of pharmaceutical mixtures in fish has been investigated at various levels of biological organization and results have not yet provided a clear direction on how management of this environmental issue should be approached. At issue is whether management of complex mixtures can be addressed by understanding effects of individual substances and predicting fish responses to complex mixtures, or whether each possible mixture combination leads to a unique organism response that must be assessed independently. The objective of the present research is to investigate whether effects of individual substances on expression of target genes (vtg and cyp1a) in zebrafish larvae can be used to predict the change in expression when fish are exposed to substance mixtures. To support observations on changes in cyp1a expression, evaluation of CYP1A enzymatic activity was assessed in parallel by evaluation of ethoxyresorufin O-deethylase (EROD) activity. Pharmaceuticals from different therapeutic groups (Table 1) were selected for this study based on their frequency of use, presence as mixtures in surface waters, and concentrations in municipal effluents (Table 2). Caffeine (Caf) is a potent stimulant of the central nervous system (Nikolau et al., 2007) and has been used as a marker for residual wastewater in the environment due to its persistance. The anti-inflammatory non-prescription drug ibuprofen (Ibu) is used as an analgesic and antipyretic and, in addition to naproxen, is one of the most abundant anti-inflammatory drugs found in municipal effluents (Miège et al., 2009; Stuer-Lauridsen et al., 2000). Carbamazepine (Cbz) is a psychiatric drug prescribed as an anticonvulsant and mood-stabilizer applied in the treatment of epilepsy, bipolar disorder, and trigeminal neuralgia (García-Morales et al., 2007). Tamoxifen (Tmx) is one of the most commonly used chemotherapeutic agents, has anti-estrogenic activity (Bergh, 2003; Osborne,

carbon receptor (AhR). The induction of CYP1A1 gene transcription by the aryl hydrocarbon begins by their binding and activating the AhR, a cytosolic protein that, on ligand binding, translocates to the nucleus and with its partner, the aryl hydrocarbon nuclear translocator, interacts with the promoter of the CYP1A1 gene (Rowlands and Gustafsson, 1997). This results in an up-regulation of transcription and a subsequent increase in CYP1A1 mRNA and enzyme levels. Consequently, this mechanism is reflected in the rated of deethylation of the substrate 7-ethoxyresorufin by cytochrome P450 (CYP) to give the product resorufin. The 7-ethoxyresorufin O-deethylase (EROD) activity is a biomarker used to determine AhR agonist exposure to certain polyhalogenated aromatic hydrocarbons (PHAHs) and polycyclic aromatic hydrocarbons (PAHs) (Bucheli and Fent, 1995; Gokøyr and Förlin, 1992) Responses of the cytochrome P450 system are used as biomarkers of oxidative stress in fish (Bucheli et al., 1995). The Induction of CYP1A by pharmaceuticals may generate reactive oxygen species (ROS) (Van der Berg et al., 1998). When a substrate is metabolized by cytochrome P450 consumes one molecule of molecular oxygen leading to an oxidized substrate plus a molecule of water as a by-product. However, for most CYPs, depending on the nature of the substrate, the reaction is "uncoupled", consuming more O2 than the metabolized substrate and producing activated oxygen or O2- (Gonzalez and Tukey, 2006). It has been demonstrated that mixtures of substances that induce cyp1a a gene that encodes for CYP1A, the induction is consistent with the addition of the individual induction activities of each substance (Hook et al., 2008; Filby et al., 2007). Induction of cyp1a and the activity of CYP1A are important because increased generation of ROS can lead to oxidative stress and damage to biomolecules, or abnormally enhance the CYP1A metabolism of endogenous substrates (Rifkind, 2006). It is also considered of greatest concern for risk assessment pharmaceuticals that bind the estrogen and induce estrogenic effects in fish such as oral contraceptives, hormone replacement therapies, motor deficits associated with menopause, hypoestrogenism, and the management of some pre- and postmenopausal symptoms (Laurenson et al., 2014). In particular, mixtures of estrogenic substances can cause effects [e.g., induction of vitellogenin genes (vtg) or vitellogenin lipoprotein (Vg) production] in fish that are reported to be equivalent to the addition of their individual activities (Filby et al., 2007; Thorpe et al., 2006, 2001). In addition, even when concentrations of individual

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Table 2 Environmental concentrations of selected pharmaceuticals. Pharmaceutical

Concentration (µg L−1)

Place of study

Reference

Caffeine

0.16 22.20 293.00

creek waste water treatment plant sewage treatment plant

Yoon et al. (2010) Gagné et al. (2006) Weigel et al. (2004)

Ibuprofen

0.01 0.05 24.60

lake river waste water treatment plant

Tixier et al. (2003) Yoon et al. (2010) Stuer-Lauridsen et al. (2000)

Carbamazepine

0.04 0.10 6.30

lake river municipal effluent

Tixier et al. (2003) Daneshvar et al. (2010) Ternes, 1988

Tamoxifen

0.002 0.21 0.37

waste water treatment plant estuary Sewage treatment plant

Lara-Martín et al. (2014) Roberts and Thomas (2006) Roberts and Thomas (2006)

2.3. Experimental design

1998; Powles et al., 1994), and has been detected in many waste water treatment plant effluents (Table 2) (Lara-Martín et al., 2014; Roberts andThomas, 2006).

Zebrafish embryos were obtained following routine procedures (Reinardy et al., 2013a, 2013b). Approximately 100 embryos were randomly selected after bulk spawning and stocked into each of multiple 500-mL glass beakers containing 400 mL of fish water for hatching. Embryos hatched in the ZRF ~72 h post fertilization (hpf), and ages of larvae used in experiments were from 72 to 96 hpf (24-h exposure), 72–120 hpf (48-h exposure), and 72–168 hpf (96-h exposure). Water quality characteristics were measured during the experiment and values were consistent with those reported for fish water, indicated above. Exposure of larvae was conducted in 300 mL covered glass beakers containing 200 mL of fish water (30–40 larvae) and the appropriate test chemical or vehicle control (0.001% EtOH). Two separate experiments were run with phenanthrene and each concentration had one beaker (n =2 experiments). Each pharmaceutical treatment and vehicle control EE2 had three replicate beakers. Larvae mortality was quantified during exposure and removed from beakers. After 96 h of exposure, live larvae were sampled and frozen at −80 °C for storage before extraction of total RNA.

2. Materials and methods 2.1. Experimental fish Zebrafish embryos were obtained from the Zebrafish Research Facility (ZRF) housed at Plymouth University, Plymouth, UK. All experiments were conducted in this facility with appropriate approval from the UK Home Office. In all experiments, zebrafish larvae were treated humanely and with regard for alleviation of suffering. Larvae were routinely bred from bulk spawning of stock fish. Water quality in the ZRF was measured and water used in these experiments (fish water) had the following characteristics: temperature 28 ± 2 °C, mean pH 7.0 ± 0.2, dissolved oxygen > 92 ± 3%, total ammonia < 0.02 mg L−1, nitrate 20 mg L−1, nitrite < 0.1 mg L−1, and electrolite composition 0.3, 0.04, 0.08, 0.4 mM for Ca, K, Mg and Na respectively. Zebrafish embryos were maintained at 27 ± 1 °C with a photoperiod of 12 h. Developing larvae were kept in 500 mL beakers with daily water change and removal of unfertilized eggs, abnormal embryos, and debris.

2.4. RNA extraction and cDNA synthesis for gene expression analyses Total RNA was extracted from sampled larvae from a single exposure beaker, following manufacturer's protocol RNeasy MiniKit for animal tissue (Qiagen, Hilden, Germany), with sonication in RLT buffer (3–5 s) for initial cell disruption. Tissues were further broken up with a QiaShredder column (Qiagen) following the manufacturer's protocol and a 15 min DNase treatment. RNA was eluted into 30 µL sterile water, and the concentration and quality of total RNA was determined by spectrophotometer (NanoDrop, ND-1000 Spectrophotometer). All samples were diluted to 100 ng µL−1 total RNA, and 800 ng were used to synthesise cDNA following the manufacturer's protocol for ImProm-II™ Reverse Transcription System (Promega, Fitchburg, WI, USA), with hexanucleotide primers and deoxynucleotide mix (Sigma-Aldrich). cDNA was synthesized under the following conditions: annealing at 25 °C, extending at 42 °C, and heat-inactivating transcriptase at 70 °C (GeneAmp® PCR System, 9700, Applied Biosystems, Waltham, MA, USA). cDNA was stored at −80 °C until q-RT-PCR gene expression analysis.

2.2. Chemicals, exposure solutions, and chemical analyses Caf, Ibu, Cbz and Tmx were purchased from Sigma Aldrich, Spain (Table 1). Test concentrations of these drugs were selected based on reported environmental concentrations from municipal effluents, wastewater treatment plant effluents, and sewage treatment plants worldwide (Table 2). Phenanthrene and 17α-ethinylestradiol (EE2) were obtained from Sigma Aldrich, UK and used as positive control substances for induction of cyp1a and vtg gene expression, respectively. Each test chemical was prepared in ethanol (1 mg mL−1 EtOH, HPLC grade). The final concentration of EtOH in each glass beaker (500 mL) was 0.001% in order to avoid a solvent effect. Chemical stocks were dissolved in EtOH to obtain a final concentration of phenanthrene (0.01, 0.05, 0.1, 0.25, 0.5, 1, and 2 mg L−1), EE2 (0.001, 0.01, 0.1, and 1 μg L−1), Caf (0.05 and 5 μM), Ibu (0.05 and 5 μM), Cbz (0.05 and 5 μM), and Tmx (0.003 and 0.3 μM). Treatment of pharmaceutical mixture corresponding to low mix (environmental concentrations) was prepared by adding separately Caf, Ibu, Cbz (0.05 μM) and Tmx (0.003 μM) to beakers, and the high mix was prepared similarly by adding Caf, Ibu, Cbz (5 μM), and Tmx (0.3 μM). All test chemicals including the mixtures were spiked to the exposure beakers containing the fish water prior to addition of the zebrafish larvae.

2.5. Selection of primers for cyp1a and vtg genes Primers for zebrafish cyp1a, vtg, and β-actin were selected based on our previous research and other published literature (Table 3). The amplicons were designed to span an intron junction and were checked to avoid secondary structure, self-annealing, complementarity, and potential hairpins by DNA calculator (Sigma-Aldrich, St Louis, MO, USA) and OligoCalc (Northwestern University, Evanston, IL, USA). After PCR amplification, amplicon size was verified by electrophoresis on a 2% agarose gel. 473

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Table 3 Primers used for qRT-PCR assessment of zebrafish cyp1a and vtg gene expression. Gene

Ref. seq. no.a

Primer

Primer Sequence (5’−3’)

Amplicon size

cyp1ab

NM131879.1

Forward Reverse

AGGACAACATCAGAGACATCACCG GATAGACAACCGCCCAGGACAGAG

174

Vtgb

NM001044897.2

Forward Reverse

ATCAGTGATGCACCTGCCCAGATTG ACGCAAGAGCTGGACAAGCTGAA

116

ß-actinc

NM131031.1.

Forward Reverse

ACACAGCCATGGATGAGGAAATCG TCACTCCCTGATGTCTGGGTCGT

138

a b c

National Center for Biotechnology Information (NCBI); Reference Sequence Number. Voelker et al. (2007). Wolińska et al. (2011).

linear regression when appropriate, for EROD activity exposed to phenanthrene dose response relations were assessed by polynomial regression. For EROD activity, significant differences between controls and treatments were determined using a one-way ANOVA (data was previously tested for normality and homogeneity of variances) followed by a multiple comparison of Dunnett's tests. Tukey's test was used to distinguish differences between treatments. Pearson's correlation was performed in order to evaluate concentration-dependent relationships and correlations between the analyzed biological responses (cyp1a gene expression and EROD activity).

2.6. Quantitative reverse transcriptase PCR (qRT-PCR) Lyophilized primers (Eurofins MWG Operon, Ebersberg, Germany) were reconstituted to 100 μmol with RNase-free water and mixed with SYBR Green JumpStart Taq ReadyMix to give a final primer concentration of 375 nmol in 20 µL total volume. Fluorescence was detected (StepOne Real-Time PCR System, Applied Biosystems) over 40 cycles, cycling conditions of 94 °C for denaturing, primer-specific annealing temperature indicated in Table 3, and extension at 72 °C. At the end of each q-PCR run, the melt curves were examined for each sample to verify single product amplification and consistency among samples. A standard curve of cDNA template (pooled template from each sample within experiment), was run on each plate to allow for withinexperiment plate normalization and determination of PCR efficiency (Reinardy et al., 2013a, 2013b).

3. Results and discussion 3.1. Fish mortality No mortality was observed in larvae exposed to control or to the solvent control (0.001% EtOH). Mortality increased with increasing phenanthrene concentration, the LC50 was 0.57 mg L−1, and all fish died at a concentration of 1 mg L−1 (Fig. 1a). This experiment was used to identify sub-lethal phenanthrene concentrations for assessment of cyp1a expression (phenanthrene was used as positive control), and lethality results were consistent with previous reports of acute toxicity of phenanthrene [LC50 =0.29 mg L−1 (Prosser et al., 2011) and 0.74 mg L−1 (Zhao et al., 2011) in adult zebrafish, and LC50 =0.49 mg L−1 (Gündel et al., 2012), and ≥0.42 mg L−1 (Butler et al., 2013) in zebrafish early life stages]. Mortality after 96 h exposure to single pharmaceuticals and mixtures were < 4%. The concentrations tested were selected to be environmentally relevant and were not expected to cause mortality. Acute toxicity of pharmaceutical substances is generally found at concentrations that are orders of magnitude higher than those applied in this experiment [e.g., reported LC50 for Caf was 3480 µM in zebrafish larvae at 72 hpf (Selderslaghs et al., 2012), in medaka (Oryzias latipes) larvae LC50 for Ibu was > 222.4 µM and for Cbz was 194.1 µM (Kim et al., 2009)].

2.7. Ethoxyresorufin O-deethylase (EROD) activity Mixed function oxidase activity was measured in sampled larvae using EROD assay initially adapted for fingerling rainbow trout (Gagné et al., 1993). Larvae were homogenized following the procedure described by Lafontaine et al. (2000) and centrifuged (15,000g for 20 min) at 2 °C to obtain the supernatant fraction. In black-walled microplates (96 flat bottom wells), 50 µL of the supernatant was added to 160 µL 7- ethoxyresorufin, and 10 µL reduced NADPH, to initiate the reaction (final volume of well 220 µL). 7- Hydroxyresorufin was detected fluorometrically every10 min for 60 min at 30 °C (kinetic microplate reader, Infinite® M200, 516 nm excitation and 600 nm emission wavelengths). Calibration was achieved through a standard curve of 7-hydroxyresorufin. Results were standardized to total protein (TP) content, measured in the supernatant using the methodology described by Bradford (1976). EROD activity was expressed as pmol min−1 mg TP−1. 2.8. Data analyses

3.2. cyp1a expression and EROD activity: phenanthrene exposure All statistical analyses were conducted with SPSS 15.0 (SPSS-IBM, Armonk, NY, USA) and a probability level of 0.05 was used as the level for statistical significance. Logistic regression was performed for the phenanthrene-induced mortality and the lethal concentration computed to cause 50% mortality (LC50) was reported with 95% confidence interval (CI). For gene expression analyses, the cycle threshold (Ct) was set to 25000 for all qRT-qPCR runs, and the Ct value was the mean of triplicate reactions. Relative quantification of cyp1a and vtg expression (Ct) was obtained by normalizing to the expression of β-actin in each sample (ΔCt). Subsequently, the difference between the mean ΔCt for the unexposed larvae (control) and larvae exposed to phenanthrene, EE2, or pharmaceuticals was computed (ΔΔCt), and the relative fold change in cyp1a and vtg expression was calculated (2−ΔΔCt) (Livak and Schmittgen, 2001). Statistical analyses of relative changes in gene expression (relative fold change) among treatments were assessed by

The expression of β-actin in zebrafish larvae did not differ significantly among vehicle controls or among the treatment chemicals tested and so the use of β-actin as a gene for normalization of target genes was considered acceptable. Expression of cyp1a significantly increased with concentration of phenanthrene, and the highest level of induction (2.5 fold) was observed at the highest exposure (5 mg L−1) to phenanthrene (Fig. 1b). The relatively low level of cyp1a induction by phenanthrene was expected as previous studies have reported phenanthrene to induce cyp1a less than other cyp1a inducers (Brzuzan et al., 2007, 2006; Levine and Oris, 1999; Wolinska et al., 2013). Phenanthrene was selected because it was expected that the pharmaceuticals in this study would likewise be relatively low cyp1a inducer. In contrast with the results in this study, Wolinska et al. (2011) reported that phenanthrene at concentrations of 0.9–9 mg L−1 did not induce cyp1a gene expression 474

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gene expression in zebrafish embryos (Incardona et al., 2006). Studies have indicated up-regulation of cyp1a gene expression in zebrafish embryos exposed to organic substances [e.g., 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) at 0.002 and 0.005 µM (Andreasen et al., 2002; Jenny et al., 2012)]. Similarly 3,4-dichloroaniline (DCA) (from 0.78 to 12.4 µM) has been shown to induce cyp1a gene expression in a concentration-dependent manner (Voelker et al., 2007). The induction of cyp1a found in the present study was not as high as the induction indicated by Jenny et al. (2012) for TCDD (135-fold induction), and is more similar to the response described after exposure to 3,4-DCA (6fold induction) (Voelker et al., 2007). In addition, it has been demonstrated that cyclopenta[c]phenanthrene (CP[c]Ph) and benzo[a]pyrene (B[a]P) induce expression of cyp1a in liver and gills (Levine and Oris, 1999), and brain (Brzuzan et al., 2007) of rainbow trout (Oncorhynchus mykiss); and in liver and gills of roach (Wolinska et al., 2013). Phenanthrene induced significantly EROD activity in zebrafish larvae and this induction increased with phenanthrene concentration (linear regression, p < 0.05; Fig. 1c). It has been suggested that induction of EROD activity is specific for PAHs (Goksøyr and Förlin, 1992). Phenanthrene has been shown to induce EROD activity in many fish species: at 0.0044 mg L−1 in cichlids (Oreochromis mossambicus) (Shailaja and Rodrigues, 2003) and at 0.05–0.48 mg L−1 in grey mullet (Liza aurata) (Oliveira et al., 2007); induction of EROD activity has been also detected in gilthead seabream (Sparus aurata) exposed at 0.02 mg L−1 and in the hybrid tilapia (Oreochromis niloticus ♀ × O. aureus ♂) when exposed to 0.05, 0.1, and 0.4 mg L−1 (Xu et al., 2009). In zebrafish, induction of EROD activity in gill and liver has been reported after exposure to benzo(a)pyrene at 0.002 mg L−1 (Jönsson et al., 2009), but there are no previous reports of effects of phenanthrene on EROD activity in zebrafish. The level of induction of EROD enzymatic activity observed in this study indicated that phenanthrene was a relatively low inducer (4.8 fold control at 0.5 mg L−1, Fig. 1C) compared to other substances previously tested in zebrafish, for example, a 13-fold induction relative to control after exposure to 1 μM ß-naphthoflavone (Jönsson et al., 2009). Results of the present study were consistent with the expectation that phenanthrene is a weak cyp1a inducer, and demonstrates a concentration-dependent response for cyp1a gene expression and EROD activity for comparison to responses from the tested pharmaceutical substances. 3.3. cyp1a expression and EROD activity: pharmaceutical exposure The effect of individual substances and their mixtures on cyp1a expression was generally consistent with the observations of effects on EROD activity. Expression of cyp1a was significantly down-regulated in larvae at all concentrations of individual substances (one way ANOVA, p < 0.05 Fig. 2a), and the effect of individual substances on EROD activity either did not differ from the control or resulted in a significant reduction in EROD activity (i.e., 0.05 µM Ibu; Fig. 2b). Regarding high mix, cyp1a expression was induced and EROD activity increased, while the low mix did not differ significantly from the controls. Gene expression of cyp1a has not been investigated previously in zebrafish larvae exposed to these pharmaceuticals; however, previous studies have reported inhibition of CYP enzymes in zebrafish after exposure to fluoxetine, ciprofloxacin, gemfibrozil, and erythromycin (Smith et al., 2012), and also in rainbow trout hepatocytes exposed to clofibrate, fenofibrate, carbamazepine, fluoxetine, and diclofenac (Laville et al., 2004). These drugs inhibited activity of CYP enzymes and may activate other metabolic pathways; it has been reported that ciprofloxacin and fluoxetine are strong CYP enzyme inhibitors in mammals (rats and humans) (McLellan et al., 1996; Murray and Murray, 2003). In contrast to inhibition of CYP, CbZ and Caf induced CYP1A enzymatic activity in the three-spined stickleback (Gasterosteus aculeatus) (Beijer et al., 2010). However, induction in the three-spined stickleback occurred at concentrations of Caf and Cbz that were much higher (200 μΜ) than concentrations found in the environment and those tested in the present

Fig. 1. Effects of phenanthrene exposure on zebrafish larvae. (a) Mortality of D. rerio larvae exposed to phenanthrene concentrations at 96 h (n =1 beaker per concentration, n =30−40 larvae per beaker). (b) Expression of cyp1a gene in larvae exposed 96 h to phenanthrene (mean ± S.E, n =2 experiments, one beaker per concentration, n =30−40 larvae per beaker). (c) Ethoxyresorufin O-deethylase (EROD) activity measured in total larvae exposed for 96 h to phenanthrene (mean ± S.E., n =3 beakers per concentration, n =30−40 larvae per beaker).

in zebrafish larvae (96 hpf). A possible explanation for the lack of induction in the Wolinska et al. (2011) study was the use of dimethyl sulfoxide (DMSO) at 0.1% v/v for preparation of stock solutions of phenanthrene, and it has been demonstrated that DMSO at 0.1% has an inhibitory effect on cyp1a expression in zebrafish larvae (Jones, 2010). Nevertheless, phenanthrene derivatives (e.g. 4-methylphenanthrene) have been shown to increase cyp1a expression almost 4-fold in the liver of roach (Rutilus rutilus) after 7 days of exposure (Wolinska et al., 2013). It has been reported that polycyclic aromatic hydrocarbons (PAHs, including phenanthrene) can activate the AhR pathway to induce cyp1a 475

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2005) and the effect of Tmx was not assessed alone. Nevertheless, in vitro studies with cultures of rainbow trout hepatocytes indicated that Tmx abolished the inhibitory action of 17β-estradiol (E2) on cyp1a expression (Navas and Segner, 2001) indicating perhaps indirect effects of Tmx on cyp1a expression. Results of the present study indicate that only the high mix led to cyp1a induction. Exposure to individual substances did not induce cyp1a, and the observed induction of the high mix cannot be predicted by the cyp1a responses to individual substances. Results of induction of cyp1a gene expression were consistent with induction of EROD activity. This consistency has been reported by Zapata-Pérez et al. (2010) in catfish (Ariopsis felis) exposed to environmental contaminants. Consistency between CYP1A protein levels and EROD activity has been reported in the European eel (Anguilla anguilla) (van der Oost et al., 1996) and numerous field studies have demonstrated a strong and significant increase in hepatic CYP1A protein levels and enzymatic activity in many species of fish from polluted environments (van der Oost et al., 2003). It has been indicated that activation of cyp1a gene in fish by inducing agents occurs via the AhR receptor; this activated gene is then transcribed to mRNA, which serves as a template for synthesis of the P450 1A1 apoprotein (Goksøy and Husøy, 1992). This study corroborates the previous statement of Goksøy and Husøy (1992) who indicated that the induction process of CYP1A could be studied at both mRNA level using cyp1a gene expression and at the enzyme level using the EROD assay. This experiment demonstrates a link between cyp1a gene expression and EROD activity. The reasons for the induction of EROD and the up-regulation of cyp1a observed in larvae exposed to the high mix of pharmaceuticals require further research probably more focused on the AhR. While results of the present study indicate induction of cyp1a and EROD activity in high mix exposures, the down-regulation of cyp1a expression with no change in EROD activity (relative to controls) for individual substance exposures has never been reported.

Fig. 2. (a) Expression of cyp1a gene and (b) Ethoxyresorufin O-deethylase (EROD) activity in D. rerio larvae exposed 96 h to caffeine (Caf), ibuprofen (Ibu), carbamazepine (Cbz) and tamoxifen (Tmx). Grey bars indicate exposure to individual substances, and hatched bars indicate exposure to mixtures (low mix: all drugs present at 0.05 μM except Tmx 0.003 μM and high mix: all drugs present at 5 μM except Tmx 0. 3 μM). *Significantly different from control (0.001% EtOH, one way ANOVA, p < 0.05). Data are means ± S.E., n =3 beakers per concentrations, n=30–40 larvae per beaker).

study. Moreover, the 8-C-hydroxylation of Caf has been identified as a marker substance for assessing the activity of CYP1A2 in rats (Kot et al., 2008), and it has been reported that Caf induces hepatic CYP1A2 in rats without association with AhR (Ayalogu et al. (1995)). The induction of CYP1A1 gene transcription is initiated by substrate binding to the AhR and interaction between the AhR-substrate complex and promoter elements of the CYP1A1 gene (Bucheli and Fent, 1995; Goksøyr and Förlin, 1992; Timme-Laragy et al., 2009). Significant induction of EROD activity relative to control organisms was observed in larvae exposed to the high mix. Some of the individual substances have been tested previously in other studies and there are some inconsistencies reported in EROD activity compared to the present results. For example, it was reported an induction of EROD activity in gill filaments of three-spined stickleback (fish) exposed to Caf at 200 μM (Beijer et al., 2010). Cbz increased CYP1A enzymatic activity in adult zebrafish at 7.5 μM (Madureira et al., 2012) and in gill filaments of three-spined stickleback at 200 μM (Beijer et al., 2010). In contrast, and in agreement with the present study, Cbz at 31.3–500 μM was shown to inhibit the basal level of EROD activity in rainbow trout hepatocytes following a concentration-dependent response (Laville et al., 2004). Results of Ibu exposure on EROD activity in Clearfin livebearer (Poeciliopsis lucida) hepatoma cells were inconsistent in that concentrations of 0.1, 1 and 20 μM did not induce EROD activity (consistent with present study in which 0.05 and 5 μM Ibu did not induce EROD) while a concentration of 10 μM induced EROD activity (Thibaut and Porte, 2008). Regarding Tmx our findings are in agreement with experiments showing that induction of CYP1A activity in fish occurred when mixed with β-flavone (Navas and Segner, 2001). Furthermore, exposure to Tmx caused a significant down-regulation of cyp1a gene expression at 0.003 μM and 0.3 μM (Fig. 2a). To the authors knowledge there are no previous studies of cyp1a gene expression in zebrafish exposed to tamoxifen in isolation. In a study on Atlantic salmon (Salmo salar) hepatocytes, it was found that Tmx and TCDD induced expression of cyp1a (Bemanian et al., 2004). This induction may be a result of the TCDD as this compound is known to be a strong cyp1a inducer (Ma and Lu, 2007; Okino et al., 2007; Silkworth et al.,

3.4. Expression of vtg Significant induction of larval zebrafish vtg gene expression was observed relative to EE2 concentration and duration of exposure/age of larvae (Fig. 3), and there was no indication of vtg induction by pharmaceuticals. A concentration-related induction of vtg by EE2 in larval zebrafish age 24–48 h post-fertilization has not been reported previously and is useful for zebrafish bioassays investigating estrogenic effects. Findings from the present study are consistent with the EE2 concentration/vtg response associations described previously (Henry et al., 2009; Roberts and Thomas, 2006) and in other studies performed in different marine and freshwater fish species (Humble et al., 2013; Meng et al., 2010; Miracle et al., 2006). A previous study indicated vtg up-regulation in adult fathead minnow (Pimephales pomelas) after 24 h of exposure to EE2 at 1 μg L−1 (Miracle et al., 2006). The expression of vtg in zebrafish larvae was significantly reduced by TMX at 0.03 and 0.3 µM compared with controls (Fig. 3). This inhibitory effect of Tmx on vtg gene expression has been demonstrated in fish hepatocytes (Bemanian et al., 2004), in adult black goby (Gobius niger) injected with 1 mg Kg−1 of Tmx (Maradona et al., 2009), and in liver of adult medaka exposed to 0.1 and 1 μM of Tmx (Yamaguchi et al., 2009). In contrast, Tmx has been reported to induce vtg gene expression in rainbow trout (Vetillard and Bailhache, 2006). An increase in VTG levels and changes in gonad histology were reported in fathead minnow embryos when exposed to Tmx at 0.02 μM (lowest observed effect concentration) (Williams et al., 2007), however in Nile tilapia, Tmx had no effect on VTG production (Leaños-Castañeda et al., 2002). In the present study, vtg expression did not differ from unexposed controls of zebrafish larvae exposed to pharmaceutical mixtures (low mix and high mix), indicating that mixtures induce different responses than can be predicted by observation of Tmx in isolation. 476

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3000 2500 2000 1500

vtg 24 h

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Fig. 3. (a) vtg gene expression in D. rerio larvae exposed 96 h to mixture of pharmaceuticals indicated in hatched lines (low mix: 0.05 µM of Caf, Ibu, Cbz and 0.003 µM of Tmx and high mix: 5 µM of Caf, Ibu, Cbz and 0.3 µM of Tmx), and to tamoxifen (Tmx) at low and high environmental concentrations. (b) Expression of vtg gene in D. rerio larvae exposed to 17αEthynylestradiol (EE2) during 24 h (black), 48 h (dark) and 96 h (grey). *Indicates significant differences with control (one way ANOVA, p < 0.05). Data are means ± S.E., n =3 beakers per concentrations, n =30−40 larvae per beaker).

antioxidant enzymes by the organism, since it could provoke LPO, DNA damage and ultimately cell death. In this sense, together with levels of CYP1A protein and gene expression, the induction of CYP1A catalytic activities may be used both for the assessment of exposure and as early-warning sign for potentially harmful effects of pharmaceuticals. The drug mixtures also affected expression of vtg in zebrafish larvae, and results suggest reduced antiestrogenic effects of Tmx.

3.5. Interpretation of actions of individual substances and complex mixtures The concepts of concentration addition and independent action are frequently employed to explain or model the effects of substances present in mixtures, but these concepts may not always be appropriate. In the present study, the effects of individual substances on the endpoints tested do not enable prediction of the effects of their mixture and this has implications on management of complex mixtures of pharmaceutical substances. It is possible that measurement of gene expression is less suitable for investigation of effects of mixtures compared to other effect endpoints. For example, evidence for concentration addition has been reported based on endpoints of immobilization of the water flea Daphnia magna and growth inhibition of green microalgae, Desmodesmus subspicatus (Cleuvers et al., 2003). It is possible that effects occurring at concentrations below detectable effect concentrations can contribute to a total effect of a mixture (e.g., concentration addition) (Fent et al., 2006), however this is not the case, in the present study individual substances caused down-regulation of target gene expression. It is possible that results of the present study can be explained by more complex synergistic effects as has been previously suggested for some rainbow trout genes that were only expressed after exposure to a chemical mixture and attributed to synergistic interactions of contaminants (Hook et al., 2008). It has been reported that mixtures of drugs present in an effluent from a drug manufacturing plant led to upregulation of cyp1a in three-spined sticklebacks (Beijer et al., 2013). However, the study of Beijer et al. (2013), and most other studies, did not investigate effects of the individual substances in isolation and it was therefore not possible to assess whether effects were additive or otherwise. Complex interactions among substances in mixtures can complicate assessment of effects (Henry and Black, 2007). For the experimental point of view, inactivation observed by single drugs exposure of the cyp1a gene and CYP1A activity may lead to bioaccumulation of pharmaceuticals and human health risk. Activation of cyp1a gene and CYP1A activity in zebrafish larvae may be traduced in the degradation and excretion of pharmaceuticals. Nevertheless, in this case, more research should be performed to determine if the oxidative stress that leads this activity is properly neutralized by

4. Conclusion Complex mixtures of pharmaceutical substances can have unpredictable effects on toxicity endpoints than those observed by the individual substances. In this study additive toxicity was not supported, and results highlight the need to evaluate mixtures of pharmaceuticals when performing an environmental risk assessment of drugs, rather than studies based on individual effects, since in the environment pharmaceutical active compounds are not found in isolation and the effects of their mixtures is poorly understood. Acknowledgements Authors would like to thank Stan McMahon for fish care and maintenance in Zebrafish Research Facility at Plymouth University. This work was conducted under the framework of the project P09-RNM5136 from Andalusian Government (Spain). G.A.M. would like to thank the financial support from Consejería de Economía, Innovación y Ciencia from the Regional Government of Andalusia (Spain), Fondos FEDER, Becas Chile (Chilean Government), and to Campus de Excelencia Internacional del Mar (CEIMAR). References Andreasen, E.A., Spitsbergen, J.M., Tanguay, R.L., Stegeman, J.J., Heideman, W., Peterson, R.E., 2002. Tissue-specific expression of AHR2, ARNT2, and CYP1A in zebrafish enbryos and larvae: effects of developmental stage and 2,3,7,8tetrachlorodibenzo-p-dioxin exposure. Toxicol. Sci. 68, 403–419. Ayalogu, E.O., Snelling, J., Lewis, D.F.V., Talwar, S., Clifford, M.N., Ioannides, C., 1995. Induction of hepatic CYP1A2 by the oral administration of caffeine to rats: lack of association with the Ah locus. Biochim. Biophys. Acta 1272 (2), 89–94. http://dx.doi. org/10.1016/0925-4439(95)00071-B.

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G.V. Aguirre-Martínez et al. Beijer, K., Abrahamson, A., Brandt, I., 2010. CYP1A inhibition in fish gill filaments: a novel assay applied on pharmaceuticals and other chemicals. Aquat. Toxicol. 96 (2), 145–150. http://dx.doi.org/10.1016/j.aquatox.2009.10.018. Beijer, K., Gao, K., Jönsson, M., Larsson, D.G.J., Brunström, B., Brandt, I., 2013. Effluent from drug manufacturing affects cytochrome P450 1 regulation and function in fish. Chemosphere 90 (3), 1149–1157. http://dx.doi.org/10.1016/j.chemosphere.2012. 09.023. Bemanian, V., Male, R., Goksøyr, A., 2004. The aryl hydrocarbon receptor-mediated disruption of vitellogenin synthesis in the fish liver: cross-talk between AHR- and ERα - signaling pathways. Comp. Hepatol. 3 (2). http://dx.doi.org/10.1186/1476-59263-2. Bergh, J., 2003. Breast-cancer prevention: is the risk-benefit ratio in favour of tamoxifen? Lancet 362, 183–184. Blasco, J., DelValls, A.T., 2008. Impact of Emergent Contaminants in the Environment: environmental Risk Assessment. Hdb. Environ. Chem. 5 (S/1), 169–188. http://dx. doi.org/10.1007/698_5_107. Bradford, M.M., 1976. A rapid and sensitive method for the quantification of microgram quantities of protein utilizing the principle of protein dye binding. Anal. Biochem. 72, 248–254. Brain, R.A., Johnson, D.J., Richard, S.M., Hanson, M.L., Sanderson, H., Lam, M.W., Young, C., Maburry, S.A., Sibley, P.K., Solomon, K.R., 2004. Microcosm evaluation of the effects of an eight pharmaceutical mixture to the aquatic macrophytes Lemna gibba and Myriophyllum sibiricum. Aquat. Toxicol. 70 (1), 23–40. http://dx.doi.org/ 10.1016/j.aquatox.2004.06.011. Brian, J.V., Harris, C.A., Scholze, M., Backhaus, T., Booy, P., Lamoree, M., Pojana, G., Jonkers, N., Runnalls, T., Bonfà, A., Marcomini, A., Sumpter, J.P., 2005. Accurate prediction of the response of freshwater fish to a mixture of estrogenic chemicals. Environ. Health Persperc. 113 (6), 721–728. Brzuzan, P., Woźny, M., Łuczyński, M.K., Góra, M., Ciesielski, S., Kuźmiński, H., 2007. Cyclopenta[c]phenanthrene induction of CYP1A in brain of rainbow trout (Oncorhynchus mykiss). Environ. Biotechnol. 3 (1), 10–14. Brzuzan, P., Jurczyk, Ł., Woźny, M., Góra, M., Łuczyński, M., K., Kuźmiński, H., Nitek, W., 2006. Molecular geometry, CYP1A gene induction and clastogenic activity of cyclopenta[c]phenanthrene in rainbow trout. Polycycl. Aromat. Compd. 26, 345–365. http://dx.doi.org/10.1080/10407780601028132. Bucheli, T.D., Fent, K., 1995. Induction of cytochrome P450 as a biomarker for environmental contamination in aquatic ecosystems. Crit. Rev. Environ. Sci. Technol. 25, 201–268. Butler, J.D., Parkerton, T.F., Letinski, D.J., Bragin, G.E., Lampi, M.A., Cooper, K.R.A., 2013. Novel passive dosing system for determining the toxicity of phenanthrene to early life stages of zebrafish. Sci. Total Environ. 463–464, 952–958 (DOI 0.1016/ j.scitotenv.2013.06.079). Christensen, A.M., Faaborg-Andersen, S., Ingerslev, F., Baun, A., 2007. Mixture and single-substance toxicity of selective serotonin reuptake inhibitors toward algae and rustaceans. Environ. Toxicol. Chem. 26 (1), 85–91. http://dx.doi.org/10.1897/06219R.1. Cleuvers, M., 2003. Aquatic ecotoxicity of pharmaceuticals including the assessment of combination effects. Toxicol. Lett. 142 (3), 185–194. http://dx.doi.org/10.1016/ S0378-4274(03)00068-7. Daneshvar, A., Svanfelt, J., Kronberg, L., Prévost, M., Weyhenmeyer, G.A., 2010. Seasonal variations in the occurrence and fate of basic and neutral pharmaceuticals in a Swedish river – lake system. Chemosphere 80 (3), 301–309. http://dx.doi.org/10. 1016/j.chemosphere.2010.03.060. Daughton, C.G., Ternes, T.A., 1999. Pharmaceuticals and personal care products in the environment: agents of subtle change? Environ. Health Persp. 107 (6), 907–938. http://dx.doi.org/10.2307/3434573. Fent, K., Weston, A., Caminada, D., 2006. Ecotoxicology of human pharmaceuticals. Aquat. Toxicol. 76 (2), 122–159. http://dx.doi.org/10.1016/j.aquatox.2005.09.009. Filby, A., Santos, E., Thorpe, K., Maack, G., Tyler, C., 2007. Gene expression profiling for understanding chemical causation of biological effects for complex mixtures: a case study on estrogens. Environ. Sci. Technol. 41 (23), 8187–8194. http://dx.doi.org/10. 1021/es071278v. Gagné, F., Baise, C., 1993. Hepatic metallotionein level and mixed function oxidase activity in fingerling rainbow trout (Oncorhynshus mykiss) after acute exposure to pulp and paper milleffluents. Water Res. 27, 1669–1682. Gagné, F., Blaise, C., André, C., 2006. Occurrence of pharmaceuticals products in a municipal effluent and toxicity to rainbow trout (Oncorhynchus mykiss) hepatocytes. Ecotox. Environ. Safe. 64 (3), 329–336. http://dx.doi.org/10.1016/j.ecoenv.2005. 04.004. García-Morales, I., Sancho Rieger, J., Gil-Nagel, A., Herranz Fernandez, J.L., 2007. Antiepileltic drugs: from a scientific evidence to clinical practice. Neurologist 13 (6 suppl 1), S20–28. http://dx.doi.org/10.1097/NRL.0b013e31815bb3b7. Gokøyr, A., Förlin, L., 1992. The cytochrome P450 system in fish, aquatic toxicology and environmental monitoring. Aquat. Toxicol. 22, 287–312. Goksøyr, A., Husøy, A.-M., 1992. The cytochrome P450 1A1 response in fish: application of immunodetection in environmental monitoring and toxicological testing. Mar. Environ. Res. 34, 147–150. Gómez, M.J., Martínez-Bueno, M.J., Lacorte, S., Fernández-Alba, A.R., Agüera, A., 2007. Pilot survey monitoring pharmaceuticals and related compounds in a sewage treatment plant located on the Mediterranean Coast. Chemosphere 66 (6), 993–1002. http://dx.doi.org/10.1016/j.chemosphere.2006.07.051. Gonzalez, F.J., Tukey, R.H., 2006. Drug metabolism. In: Brunton, L.L., Lazo, J.S., Parker, K.L. (Eds.), Goodman and Gilman's The Pharmacologic Basis of Therapeutics, 11th edition. McGraw Hill, New York, pp. 71–91. Gros, M., Petrović, M., Barceló, D., 2007. Wastewater treatment plants as a pathway for aquatic contamination by pharmaceuticals in the Ebro river basin (northeast Spain).

Environ. Toxicol. Chem. 26 (8), 1553–1562. http://dx.doi.org/10.1897/06-495R.1. Gündel, U., Kalkhof, S., Zitzkat, D., vonBergen, M., Altenburger, R., Küster, E., 2012. Concentration–response concept in ecotoxicoproteomics: effects of different of different phenanthrene concentrations to the zebrafish (Danio rerio) embryo proteome. Ecotoxicol. Environ. Saf. 76 (2), 11–22. http://dx.doi.org/10.1016/j. ecoenv.2011.10.010. Halling-Sørensen, B., Nors Nielsen, S., Lanzky, P.F., Ingerslev, F., Holten Lützheft, H.C., Jørgensen, S.E., 1998. Ocurrence, fate and effects of pharmaceuticals substances in the environment - A review. Chemosphere 36 (2), 357–393. http://dx.doi.org/10. 1016/S0045-6535(97)00354-8. Henry, T.B., Black, M.C., 2007. Acute toxicity of mixtures of selective serotonin reuptake inhibitors to Ceriodaphnia dubia. Environ. Toxicol. Chem. 26 (8), 1751–1755. http:// dx.doi.org/10.1897/06-265R.1. Henry, T.B., McPherson, J.T., Rogers, E.D., Heah, T.P., Hawkins, S.A., Layton, A.C., Sayler, G.S., 2009. Changes in the relative expression pattern of multiple vitellogenin genes in adult male and larval zebrafish exposed to exogenous estrogens. Comp. Biochem. Phys. A. 154 (1), 119–126. http://dx.doi.org/10.1016/j.cbpa.2009.05.009. Hook, S.E., Skillman, A.D., Gopalan, B., Small, J.A., Schultz, I.R., 2008. Gene expression profiles in rainbow trout, Onchorynchus mykiss, exposed to a simple chemical mixture. Toxicol. Sci. 102 (1), 42–60. http://dx.doi.org/10.1093/toxsci/kfm293. Humble, J.L., Hands, E., Saaristo, M., Lindström, K., Lehtonen, K.K., Diaz de Cerio, O., Cancio, I., Wilson, G., Craft, J.A., 2013. Characterization of genes transcriptionally upregulated in the liver of sand goby (Pomatoschistus minutus) by 17aethinyloestradiol: Identification of distinct vitellogenin and zona radiata protein transcripts. Chemosphere 90 (11), 2722–2729. http://dx.doi.org/10.1016/j. chemosphere.2012.11.053. Incardona, J.P., Day, H.L., Collier, T.K., Scholz, N.L., 2006. Developmental toxicity of 4ring polycyclic aromatic hydrocarbons in zebrafish is differentially dependent on AH receptor isoforms and hepatic cytochrome P4501A metabolism. Toxicol. Appl. Pharm. 217 (3), 308–321. http://dx.doi.org/10.1016/j.taap.2006.09.018. Jenny, M.J., Aluru, N., Hahn, M.E., 2012. Effects of short-term exposure to 2,3,7,8tetrachlorodibenzo-p-dioxin on microRNA expression in zebrafish embryos. Toxicol. Appl. Pharm. 264, 262–273. Jones, O.A.H., Voulvoulis, N., Lester, J.N., 2005. Human pharmaceuticals in wastewater treatment processes. Crit. Rev. Environ. Sci. Tec. 35, 401–427. Jones, S.H., 2010. Xenobiotic metabolism and zebrafish (Danio rerio) larvae. Thesis for the defgree of Doctor of Philosophy. School of Biosciences, College of life and Environmental Sciences. University of Birminghampp. 154. Jönsson, M.E., Brunström, B., Brand, I., 2009. The zebrafish gill model: induction of CYP1A, EROD and PAH adduct formation. Aquat. Toxicol. 91, 62–70. http://dx.doi. org/10.1016/j.aquatox.2008.10.010. Ketan, S.K., Collins, T.J., 2007. Human pharmaceuticals in the aquatic environment: a challenge to green chemistry. Chem. Rev. 107, 2319–2364. http://dx.doi.org/10. 1021/cr020441w. Kim, J.W., Ishibashi, H., Yamauchi, R., Ichikawa, N., Takao, Y., Hirano, M., Koga, M., Arizono, K., 2009. Acute toxicity of pharmaceutical and personal care products on freshwater crustacean (Thamnocephalus platyurus) and fish (Oryzias latipes). J. Toxicol. Sci. 34 (2), 227–232. http://dx.doi.org/10.2131/jts.34.227. Kot, M., Daniel, W., 2008. Caffeine as a marker substrate for testing cytochrome P450 activity in human and rat. Pharmacol. Rep. 60 (6), 789–797. Lafontaine, Y., Gagné, F., Blaise, C., Costan, G., Gagnon, P., Chan, H.M., 2000. Biomarkers in Zebra mussels (Dreissena polymorpha) for the assessment and monitoring of water quality of the St. Lawrence River (Canada). Aquat. Toxicol. 50, 51–70. Lara-Martín, P., González-Mazo, E., Petrovic, M., Barceló, D., Brownawell, B., 2014. Occurrence, distribution and partitioning of nonionic surfactants and pharmaceuticals in the urbanized Long Island Sound Estuary (NY). Mar. Pollut. Bull. 85 (2), 710–719. http://dx.doi.org/10.1016/j.marpolbul.2014.01.022. Laurenson, J.P., Bloom, R.A., Page, S., Sadrieh, N., 2014. Ethinyl estradiol and other human pharmaceutical estrogens in the aquatic environment: a review of recent risk assessment data. AAPS J. 16 (2), 299–301. http://dx.doi.org/10.1208/s12248-0149561-3. Laville, N., Ait-Aissa, S., Gomez, E., Casellas, C., Porcher, J.M., 2004. Effects of human pharmaceuticals on cytotoxicity, EROD activity and ROS production in fish hepatocytes. Toxicology 196 (1–2), 41–55. http://dx.doi.org/10.1016/j.tox.2003.11. 002. Leaños-Castañeda, O., Van Der Kraak, G., Lister, A., 2002. Simá-Alvarez, R., GoldBouchot, G. o, p’-DDT induction of vitellogenesis and its inhibition by tamoxifen in Nile tilapia (Oreochromis niloticus). Mar. Environ. Res. 54 (3–5), 703–707. http://dx. doi.org/10.1016/S0141-1136(02)00168-X. Levine, S.L., Oris, J.T., 1999. CYP1A expression in liver and gill of rainbow trout following waterborne exposure: implications for biomarker determination. Aquat. Toxicol. 46 (3–4), 279–287. http://dx.doi.org/10.1016/S0166-445X(98)00124-6. Livak, K.J., Schmittgen, T.D., 2001. Analysis of relative gene expression data using realtime quantitative PCR and the 2(T) (−Delta Delta C) method. Methods 25, 402–408. http://dx.doi.org/10.1006/meth.2001.1262. Ma, Q., Lu, A.Y., 2007. CYP1A induction and human risk assessment: an evolving tale of in iitro and in vivo studies. Drug. Metab. Dispos. 35 (7), 1009–1016. http://dx.doi. org/10.1124/dmd.107.015826. Madureira, T.V., Rocha, M.J., Cruzeiro, C., Rodrigues, I., Monteiro, R.A.F., Rocha, E., 2012. The toxicity potential of pharmaceuticals found in the Douro River estuary (Portugal): evaluation of impacts on fish liver, by histopathology, stereology, vitellogenin and CYP1A immunohistochemistry, after sub-acute exposures of the zebrafish model. Environ. Toxicol. Pharm. 34 (1), 34–35. http://dx.doi.org/10.1016/ j.etap.2012.02.007. Maradona, F., Batti, S., Marino, M., Mita, D.G., Carnevalia, O., 2009. Tamoxifen as an Emerging Endocrine Disruptor Effects on Fish Reproduction and detoxification Target

478

Ecotoxicology and Environmental Safety 142 (2017) 471–479

G.V. Aguirre-Martínez et al.

phenanthrene in fish and its implications for coastal waters. Estuar. Coast. Shelf S. 56, 1107–1110. http://dx.doi.org/10.1016/S0272-7714(02)00311-6. Silkworth, J.B., Koganti, A., Illouz, K., Possolo, A., Zhao, M., Hamilton, S.B., 2005. Comparison of TCDD and PCB CYP1A induction sensitivities in fresh hepatocytes from human donors, Sprague-Dawley rats, and Rhesus monkeys and HepG2 cells. Toxicol. Sci. 87 (2), 508–519. http://dx.doi.org/10.1093/toxsci/kfi261. Smith, E.M., Iftikar, F.I., Higgins, S., Irshad, A., Jandoc, R., Lee, M., Wilson, J.Y., 2012. In vitro inhibition of cytochrome P450-mediated reactions by gemfibrozil, erythromycin, ciprofloxacin and fluoxetine in fish liver microsomes. Aquat. Toxicol. 109, 259–266. http://dx.doi.org/10.1016/j.aquatox.2011.08.022. Stuer-Lauridsen, F., Birkved, M., Hansen, L.P., Holten Lützhøftb, H.C., Halling-Sørensen, B., 2000. Environmental risk assessment of human pharmaceuticals in Denmark after normal therapeutic use. Chemosphere 40, 783–793. Ternes, T.A., 1998. Occurrence of drugs in German sewage treatment plants and rivers. Water Res. 32 (11), 3245–3260. http://dx.doi.org/10.1016/S0043-1354(98) 00099-2. Thibaut, R., Porte, C., 2008. Effects of fibrates, anti-inflammatory drugs antidepressants in the fish hepatoma cell line PHLC-1: cytotoxicity and interactions with cytochrome P450 1A. Toxicol. Vitr. 22, 1128–1135. Thomas, P.M., Foster, G.D., 2004. Determination of nonsteroidal anti-inflammatory drugs, caffeine, and triclosan in wastewater by gas chromatography-mass spectrometry. J. Environ. Sci. Heal. A. 39 (8), 1969–1978. http://dx.doi.org/10. 1081/ESE-120039368. Thorpe, K.L., Gross-Sorokin, M., Johnson, I., Brighty, G., Tyler, C.R., 2006. An assessment of the model of concentration addition for predicting the estrogenic activity of chemical mixtures in wastewater treatment works effluents. Environ. Health Persp. 114 (1), 90–97. http://dx.doi.org/10.1289/ehp.8059. Thorpe, K.L., Hutchinson, T.H., Hetheridge, M.J., Scholze, M., Sumpter, J.P., Thyler, C.R., 2001. Assessing the biological potency of binary mixture of environmental extrogens using vitelogenin induction in juvenile rainbow trout (Onchorhynchus mykiss). Environ. Sci. Technol. 35, 2476–2481. http://dx.doi.org/10.1021/es001767u. Timme-Laragy, A.R., Cockman, C.J., Matson, C.W., Di Giulio, R.T., 2009. Synergistic induction of AHR regulated genes in developmental toxicity from co-exposure to two model PAHs in zebrafish. Aquat. Toxicol. 85, 241–250. http://dx.doi.org/10.1016/j. aquatox.2007.09.005. Tixier, C., Singer, H.P., Ollers, S., Müller, S.R., 2003. Occurrence and fate of carbamazepine, clorifibric acid, diclofenac, ibuprofen, ketoprofen, and naproxen in surface waters. Environ. Sci. Technol. 37, 1061–1068. Van der Berg, M., Birnbaum, L., Bosveld, A.T., Brunström, B., Cook, P., et al., 1998. Toxic Equivalency Factors (TEFs) for PCBs, PCDDs, PCDFs for Humans and Wildlife. Environ. Health Perspect. 106 (12), 775–792. Vetillard, A., Bailhache, T., 2006. Effects of 4-n-nonylphenol and tamoxifen on salmon gonadotropin-releasing hormone, estrogen receptor, and vitellogenin gene expression in juvenile rainbow trout. Toxicol. Sci. 92 (2), 537–544. http://dx.doi.org/10.1093/ toxsci/kfl015. Voelker, D., Vess, C., Tillmann, M., Nagel, R., Otto, G.W., Geisler, R., Schirmer, K., Scholz, S., 2007. Differential gene expression as a toxicant-sensitive endpoint in zebrafish embryos and larvae. Aquat. Toxicol. 81 (4), 355–364. http://dx.doi.org/10.1016/j. aquatox.2006.12.01. Weigel, S., Berguer, U., Jensen, E., Kallenborn, R., Thorensen, H., Hühnerfuss, H., 2004. Determination of selected pharmaceuticals and caffeine in sewage and seawater from Tromsø/Norway with emphasis on ibuprofen and its metabolites. Chemosphere 56, 583–592. Williams, T.D., Caunter, J.E., Lillicrap, A.D., Hutchinson, T.H., Gillings, E.G., Duffell, S., 2007. Evaluation of the reproductive effects of tamoxifen citrate in partial and full life-cycle studies using fathead minnows (Pimephales promelas). Environ. Toxicol. Chem. 26 (4), 695–707. http://dx.doi.org/10.1897/05-646R1.1. Wolinska, L., Brzuzan, P., Woźny, M., Łuczyński, M.K., Góra, M., 2013. CYP1A Expression in liver and gills of roach (Rutilus rutilus) after waterborne exposure to two phenanthrene derivatives, 1-Methylphenanthrene and 4-Methylphenanthrene. Environ. Toxicol. Chem. 32 (7), 1604–1610. http://dx.doi.org/10.1002/etc.2225. Wolinska, L., Brzuzan, P., Woźny, M., Góra, M., Łuczyński, M.K., Podlasz, P., Kolwicz, S., Piasecka, A., 2011. Preliminary study on adverse effects of phenanthrene and its methyl and phenyl derivatives in larval zebrafish, Danio rerio. Environ. Biotechnol. 7 (1), 26–33. Xu, W., Li, Y., Wu, Q., Wang, S., Zheng, H., Liu, W., 2009. Effects of phenanthrene on hepatic enzymatic activities in tilapia (Oreochromis niloticus ♀ × O. aureus ♂). J. Environ. Sci. 21, 854–857. http://dx.doi.org/10.1016/S1001-0742(08)62352-9. Yoon, Y., Ryu, J., Oh, J., Choi, B.G., Snyder, S.A., 2010. Occurrence of endocrine disrupting compounds, pharmaceuticals, and personal care products in the Han River (Seoul, South Korea). Sci. Total Environ. 408 (3), 636–643. http://dx.doi.org/10. 1016/j.scitotenv.2009.10.049. Zapata-Pérez, O., Sanchez-Teyer, L.F., Perez-Nuñez, M.T., Arroyo-Herrera, A.L., QuirozMoreno, A., Albores-Medina, A., 2010. Response of CYP1A gene expression in fish liver of catfish (Ariopsis felis) from Gulf of Mexico and their relationship with the genetic variability. J. Environ. Sci. Heal. A. 45 (1), 62–67. http://dx.doi.org/10. 1080/10934520903388830.

Genes. Ann. NY Acad. Sci. 1163, 457–459. http://dx.doi.org/10.1111/j.1749-6632. 2008.03653.x. McLellan, R.A., Drobitch, R.K., Monshouwer, M., Renton, K.W., 1996. Fluoroquinolone antibiotics inhibit cytochrome P450-mediated microsomal drug metabolism in rat and human. Drug. Metab. Dispos. 24 (10), 1134–1138. Meng, X., Bartholomew, C., Craft, J.A., 2010. Differential expression of vitellogenin and oestrogen receptor genes in the liver of zebrafish Danio rerio. Anal. Bioanal. Chem. 396 (2), 625–630. http://dx.doi.org/10.1007/s00216-009-3112-2. Miège, C., Choubert, J.M., Ribeiro, L., Eusèbe, M., Coquery, M., 2009. Fate of pharmaceuticals and personal care products in wastewater treatment plants Conception of a database and first results. Environ. Pollut. 157 (5), 1721–1726. http://dx.doi.org/10.1016/j.envpol.2008.11.045. Miracle, A., Ankley, G., Lattier, D., 2006. Expression of two vitellogenin genes (vg1 and vg3) in fathead minnow (Pimephales promelas) liver in response to exposure to steroidal estrogens and androgens. Ecotox. Environ. Safe. 63 (3), 337–342. http://dx. doi.org/10.1016/j.ecoenv.2005.12.002. Murray, M., Murray, K., 2003. Mechanism-based inhibition of CYP activities in rat liver by fluoxetine and structurally similar alkylamines. Xenobiotica 33 (10), 973–987. http://dx.doi.org/10.1080/00498250310001602748. Nakada, N., Tanishima, T., Shinohara, H., Kiri, K., Takada, H., 2006. Pharmaceutical chemicals and endocrine disrupters in municipal wastewater in Tokyo and their removal during activated sludge treatment. Water Res. 40 (17), 3297–3303. http:// dx.doi.org/10.1016/j.watres.2006.06.039. Navas, J., Segner, H., 2001. Estrogen-mediated suppression of cytochrome P4501A (CYP1A) expression in rainbow trout hepatocytes: role of estrogen receptor. Chem. Biol. Interact. 138 (3), 285–298. http://dx.doi.org/10.1016/S0009-2797(01) 00280-0. Nikolau, A., Meric, S., Fatta, D., 2007. Occurrence patterns of pharmaceuticals in water and wastewater environments. Anal. Bioanal. Chem. 387 (4), 1225–1234. http://dx. doi.org/10.1007/s00216-006-1035-8. Okino, S.T., Quattrochi, L.C., Pookot, D., Iwahashi, M., Dahiya, R.A., 2007. Dioxinresponsive enhancer 3′ of the human CYP1A2 Gene. Mol. Pharmacol. 72, 1457–1465. http://dx.doi.org/10.1124/mol.107.039826. Oliveira, M., Pacheco, M., Santos, M.A., 2007. Cytochrome P4501A, genotoxic and stress responses in golden grey mullet (Liza aurata) following short-term exposure to phenanthrene. Chemosphere 66, 1284–1291. http://dx.doi.org/10.1016/j. chemosphere.2006.07.024. van der Oost, R., Beyer, J., Vermeulen, N.P.E., 2003. Fish bioaccumulation and biomarkers in environmental risk assessment: a review. Environ. Toxicol. Phar. 13, 57–149. http://dx.doi.org/10.1016/S1382-6689(02)00126-6. van der Oost, R., Goksøyr, A., Celander, M., Heida, H., Vermeulen, N.P.E., 1996. Biomonitoring of aquatic pollution with feral eel (Anguilla anguilla) II. biomarkers: pollution-induced biochemical responses. Aquat. Toxicol. 36, 189–222. http://dx. doi.org/10.1016/S0166-445X(96)00802-8. Osborne, C.K., 1998. Steroid hormone receptors in breast cancer management. Breast Cancer Res. Tr. 51 (3), 227–238. Powles, T.J., Jones, A.L., Ashley, S.E., O'Brien, M.E.R., Tidy, V.A., Treleavan, J., Cosgrove, D., Nash, A.G., Sacks, N., Baum, M., McKinna, J.A., Davey, J.B., 1994. The Royal Marsden Hospital pilot tamoxifen chemoprevention trial. Breast Cancer Res. Tr. 31, 73–82. Prosser, C.M., Unger, M.A., Vogelbein, W.K., 2011. Multistressor interactions in the zebrafish (Danio rerio): concurrent phenanthrene exposure and Mycobacterium marinum infection. Aquat. Toxicol. 102, 177–185. http://dx.doi.org/10.1016/j. aquatox.2011.01.011. Reinardy, H.C., Syrett, J.R., Jeffree, R.A., Henry, T.B., Jha, A.N., 2013b. Cobalt-induced genotoxicity in male zebrafish (Danio rerio), with implications for reproduction and expression of DNA repair genes. Aquat. Toxicol. 126 (15), 224–230. http://dx.doi. org/10.1016/j.aquatox.2012.11.007. Reinardy, H.C., Scarlett, A.G., Henry, T.B., West, C.E., Hewitt, L.M., Frank, R.A., Rowland, S.J., 2013a. Aromatic naphthenic acids in oil sands process-affected water, fully resolved by GCxGC-MS, weakly induce the gene for vitellogenin production in zebrafish (Danio rerio) larvae. Environ. Sci. Technol. 47 (12), 6614–6620. http://dx. doi.org/10.1021/es304799m. Rowlands, J.C., Gustafsson, J.A., 1997. Aryl hydrocarbon receptor-mediated signal transduction 1997. Crit. Rev. Toxicol. 27 (2), 109–134. http://dx.doi.org/10.3109/ 10408449709021615. Rifkind, A.B., 2006. CYP1A in TCDD toxicity and in physiology-with particular reference to CYP dependent arachidonic acid metabolisms and other endogenous substrates. Drug Metab. Rev. 38 (1–2), 291–335. http://dx.doi.org/10.1080/ 03602530600570107. Roberts, P.H., Thomas, K.V., 2006. The occurrence of selected pharmaceuticals in wastewaters effluent and surface waters of lower Tyne catchment. Sci. Total Environ. 356 (1–3), 143–153. http://dx.doi.org/10.1016/j.scitotenv.2005.04.03. Selderslaghs, I.W.T., Blustb, R., Witters, H.E., 2012. Feasibility study of the zebrafish assay as an alternative method to screen for developmental toxicity and embryotoxicity using a training set of 27 compounds. Reprod. Toxicol. 33, 142–154. Shailaja, M.S., Rodrigues, A., 2003. Nitrite-induced enhancement of toxicity of

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