Steroid androgen 17α-methyltestosterone induces malformations and biochemical alterations in zebrafish embryos

Environmental Toxicology and Pharmacology 44 (2016) 107–113

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Steroid androgen 17␣-methyltestosterone induces malformations and biochemical alterations in zebrafish embryos Carla Letícia Gediel Rivero-Wendt a,b , Rhaul Oliveira b , Marta Sofia Monteiro c , Inês Domingues c,∗ , Amadeu Mortágua Velho Maia Soares c,d , Cesar Koppe Grisolia b a Departament of Biology, University Anhanguera—Uniderp, Campus Agrárias, R. Alexandre Herculano, 1400, Taquaral Bosque, CEP 79035-470 Campo Grande, MS, Brazil b Department of Genetics and Morphology, Institute of Biological Sciences, University of Brasília, Asa Norte, CEP 70910-900 Brasília, DF, Brazil c Department of Biology & CESAM, University of Aveiro, Campus de Santiago, 3810-193 Aveiro, Portugal d Programa de Pós-Graduac¸ão em Produc¸ão Vegetal, Universidade Federal do Tocantins, Campus de Gurupi. Rua Badejós, Zona Rural, Cx. Postal 66, CEP 77402-970 Gurupi, TO, Brazil

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Article history: Received 22 February 2016 Received in revised form 20 April 2016 Accepted 23 April 2016 Available online 27 April 2016 Keywords: Biomarkers Endocrine disruptor Fish embryo test Oxidative stress Vitellogenin

a b s t r a c t The synthetic androgen 17␣-methyltestosterone is widely used in fish aquaculture for sex reversion of female individuals. Little is known about the amount of MT residues reaching the aquatic environment and further impacts in non-target organisms, including fish early-life stages. Thus, in this work, zebrafish embryos were exposed to two forms of 17␣-methyltestosterone: the pure compound (MT) and a formulation commonly used in Brazil (cMT). For MT, a 96 h-LC50 of 10.09 mg/l was calculated. MT also affected embryo development inducing tail malformations, edemas, abnormal development of the head, and hatching delay. At biochemical level MT inhibited vitellogenin (VTG) and inhibited cholinesterase and lactate dehydrogenase. cMT elicited similar patterns of toxicity as the pure compound (MT). Effects reported in this study suggest a potential environmental risk of MT, especially since the VTG effects occurred at environmental relevant concentrations (0.004 mg/l). © 2016 Elsevier B.V. All rights reserved.

1. Introduction The synthetic androgen 17␣-methyltestosterone (MT) is worldwide known for its use in fish hatcheries to induce male monosex cultures (Johnstone et al., 1983). Male individuals are of greater economic interest since they have higher growth rate given that high energy losses associated with female gonadal development and egg production are avoided (Mlalila et al., 2015). The use of MT is a regular procedure in the production of Oreochromis niloticus (tilapia) where the compound is added to the diet of juveniles to induce sex reversion in genetically defined females (Clemens and Inslee, 1968; Pandian and Sheela, 1995). Tilapia has become the second fish species in aquaculture production, with a worldwide estimated production exceeding 4.85 million tonnes in 2014 which is mainly concentrated in Asian and Latin-American countries (FAO,

∗ Corresponding author. E-mail addresses: [email protected] (C.L.G. Rivero-Wendt), [email protected] (R. Oliveira), [email protected] (M.S. Monteiro), [email protected] (I. Domingues), [email protected] (A.M.V.M. Soares), [email protected] (C.K. Grisolia). http://dx.doi.org/10.1016/j.etap.2016.04.014 1382-6689/© 2016 Elsevier B.V. All rights reserved.

2015). However, the intensive use of MT in tilapia aquaculture still poses concerns regarding the risks to consumers and environmental health (Mlalila et al., 2015). Steroid hormones such as MT are quickly adsorbed into the sediments due to their hydrophobic nature. Additionally, biotransformation and photodegradation processes also contribute to a relatively rapid disappearance of MT from the water column although these processes are dependent on several factors such as pH and organic content of the medium (Mlalila et al., 2015). Nonetheless, studies reporting environmental concentrations of MT both in water and sediments are scarce, preventing an accurate prediction of environmental risk. Megbowon and Mojekwu (2014) reported a MT accumulation in the sediment of tilapia ponds reaching 2–6 mg/kg after 28 days of the onset of fish feeding with MT-impregnated food, while Barbosa et al. (2013) analysed water from tilapia farming ponds and surrounding aquatic channels in Thailand, finding concentrations above 60 ␮g/l of MT in 6 out of 26 samples analysed. The effects of androgenic chemicals on wildlife have received increased attention in last decades with some studies accounting for the direct contribution of these compounds to the masculinization of aquatic organisms such as the mosquito fish (Gambusia affinis holbrooki) (Cody and Bortone, 1997; Parks et al., 2001).

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While the effects of MT on the target species are known, possible adverse effects in non-target species are not fully understood. Several authors have demonstrated the adverse effects of MT exposure on the sexual development and reproduction of juvenile and adult fish species, by showing declines in fertility and fecundity, development of male secondary sex characteristics and inhibition of gonadal development. For instance, these effects could be seen after chronic exposure of fathead minnow (Pimephales promelas) (Ankley et al., 2001), medaka (Oryzias latipes) (Kang et al., 2008), Astyanax bimaculatus (a Brazilian native species) and nile tilapia (O. niloticus) (Rivero-Wendt et al., 2013) to MT as well as after a full life cycle test also with medaka (Seki et al., 2004). MT effects on the reproductive system were also reported in invertebrates such as the snail Biomphalaria glabrata (Rivero-Wendt et al., 2014) and the tomato moth Lacanobia oleracea (Kirkbride-Smith et al., 2001). Less attention has been paid to the effects of MT during early life stages of organisms. In view of this, the fish embryo toxicity (FET) test as become a useful tool to assess chemical effects not only at lethal level but also at developmental level. Furthermore, biochemical biomarkers can be added as endpoints to this test, providing information at sub-individual level and contributing to a better understanding of the mode of action of the compound. In this work different biomarkers implicated in key physiological processes were chosen: phospholipoprotein vitellogenin (VTG), lactate dehydrogenase (LDH), cholinesterase (ChE), glutathione-Stransferase (GST) and catalase (CAT). VTG is the precursor of the egg yolk, that is produced through activation of the estrogen receptor and controlled through the hypothalamic–pituitary–gonadal axis (Ankley et al., 2005). The vitellogenin production in male and juvenile fish is a widely accepted biomarker of estrogenic exposure and a key endpoint for the evaluation of endocrine disruptors in field and laboratory studies (Hutchinson et al., 2006). LDH is a key enzyme in the anaerobic pathway of energy production, responsible for converting pyruvate to lactate in the absence of oxygen and is activated in conditions of chemical stress when high levels of energy are required (De Coen et al., 2001). The enzyme ChE has an important role in the neurotransmission mediated by the transmitter acetylcholine and thus is essential to the maintenance of nerve function (Olsen et al., 2001). GSTs are a family of proteins implicated in the phase II of the detoxification process, responsible for the biotransformation of xenobiotic compounds and endogenous substances (Hyne and Maher, 2003). CAT is an enzyme belonging to the anti-oxidant defence system responsible for the breakdown hydrogen peroxide, a reactive oxygen species (Oruc and Uner, 2000). In this study, we aim to understand the effects of MT in the early development of fish, using zebrafish (Danio rerio) embryos as model species. Therefore, the main objectives of this work were: (i) to evaluate MT effects on embryo survival and development, (ii) to evaluate biochemical responses after sub-lethal exposure to MT through the analysis of biomarkers of endocrine disruption (VTG), energetic metabolism (LDH), neurological stress (ChE) and oxidative stress (GST and CAT) and iii) to compare the toxicity of MT (pure compound) with a formulation widely used in tilapia aquaculture in Brazil (herein referred as commercial MT (cMT)).

Brazil). Stock solutions were prepared by dissolving 13 mg of each compound in 100 ␮l of dimethylsulphoxide (DMSO) and 99.9 ml of water. Test solutions were then prepared by diluting the stock solution in culture water (see below). A solvent control (containing 100 ␮l/l of DMSO, which corresponds to the concentration of solvent used in the highest MT concentration tested) was included in all assays performed. 2.2. Fish embryo toxicity test Zebrafish were kept in a ZebTEC (Tecniplast, Buguggiate, Italy) recirculating system. Culture water was obtained through reverse osmosis and activated carbon filtration of tap water, complemented with 0.34 mg/l of salt (“Instant Ocean Synthetic Sea Salt”, Spectrum Brands, USA) and automatically adjusted for pH and conductivity. Water temperature was 26.0 ± 1 ◦ C, conductivity 750 ± 50 ␮S, pH 7.5 ± 0.5 and dissolved oxygen equal or above 95% saturation. A 12:12 h photoperiod cycle was maintained. The adult fish were fed twice a day with artificial diet (ZM-400 fish food; Zebrafish Management Ltd, Winchester, UK) and brine shrimp. Eggs were obtained by breeding of fish in aquaria with marbles in the bottom to protect eggs from predation by adults. The day prior to breeding, males and females were separated by placing a barrier in the holding container until the next morning. Early morning, barriers were removed and fish allowed to breed. After removal of marbles, eggs were collected, rinsed in water and checked under a stereomicroscope (Stereoscopic Zoom Microscope-SMZ 1500, Nikon Corporation). Eggs with cleavage irregularities, injuries or other kind of malformations were discarded. Assays were based on the OECD guideline on Fish Embryo Toxicity Test (OECD, 2013). Ten eggs were used per treatment (3 replicates per treatment) and distributed in 24well microplates. Eggs were placed in each well individually with 2 ml of test solution and exposed to MT or cMT nominal concentrations of 0 (negative control), 5, 7, 9, 11 and 13 mg/l plus the control solvent. As cMT formulation contains 90% MT, the nominal concentrations of cMT were considered hereafter as 0, 4.5, 6.3, 8.1, 9.9 and 11.7 mg/l. Tests ran for 120 h in controlled conditions of temperature (26.0 ± 1 ◦ C) and photoperiod (12 h light: 12 h dark). Embryos were daily observed with the help of stereomicroscopy using a magnification of 70 for non-hatched embryos and 40 for hatched embryos. In the embryo phase, the following parameters were evaluated: egg coagulation, otolith formation, eye and body pigmentation, somite formation, heart-beat, tail circulation, detachment of the tail-bud from the yolk sac and hatching. After hatching, the following parameters were evaluated: eye malformation, cardiac and yolk-sac edema, tail malformation, and mortality. 2.3. Biomarkers test For biomarker analyses, newly fertilized zebrafish embryos were exposed during 96 hours to the nominal concentrations of 0, 0.004, 0.023, 0.139, 0.833, and 5 mg/l of MT or 0, 0.004, 0.021, 0.125, 0.75 and 4.5 mg/l of cMT. Ten embryos per replicate were exposed in Petri dishes with 15 ml of test solution and at least 13 replicates per treatment were used. At the end of the test, pools of 10 organisms were snap frozen in microtubes in liquid nitrogen (at least 5 pools per treatment for VTG analysis and 8 pools per treatment for other biomarkers). Samples were then kept at −80 ◦ C until analysis.

2. Material and methods 2.1. Chemicals and test solutions The compound 17␣-methyltestosterone (99.9% of purity, Empirical Formula: C20 H30 O2 , CAS: 58-18-4) was purchased from Sigma Aldrich (Co., St. Louis, MO). The commercial formulation of MT (cMT, 90% purity) was purchased from Bioativa (Paraná,

2.3.1. Vitellogenin like-proteins quantification Samples were homogenized through sonication in homogenization buffer (125 mM NaCl, 25 mM Tris-HCl, 5 mM EDTA and 1 mM dithiothreitol at pH 8; ratio 1 ml to 200 g of tissue) and then centrifuged at 12000g for 20 min at 4 ◦ C. The supernatant was then used to quantify total protein content and VTG-like proteins levels. VTG was then determined by the indirect alkali-labile

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phosphate method following the protocol presented in Gagné and Blaise (2000) with some modifications (Hallgren et al., 2009). Briefly, 100 ␮l of the supernatant were mixed with 54 ␮l of acetone (35% of final volume) during 5–10 min at room temperature and then mixed with a vortex at least three times and then centrifuged at 10000g for 5 min. After acetone removal, 50–100 ␮l (depending of pellet size) of 1 M NaOH were added to samples that were then maintained for 90 min at 70 ◦ C (Hallgren et al., 2009) to allow hydrolysis of bound phosphate (Gagné and Blaise, 2000). The levels of free phosphates were determined in the aqueous phase according to the phosphomolybdenium method (Stanton, 1968). Results are expressed as ␮g PO4 /mg protein. 2.3.2. Lactate dehydrogenase, ChE, GST and CAT determinations Samples were defrosted on ice, homogenized in phosphate buffer (0.1 M, pH 7.4) using a sonicator (KIKA Labortechnik U2005Control TM) and centrifuged for 20 min at 11500g in order to isolate the post-mitochondrial supernatant (PMS). LDH activity was measured in the PMS at 340 nm, by continuously monitoring (during 5 min) the decrease in absorbance due to the oxidation of NADH, following the methodology described by Vassault (1983) with the modifications introduced by Diamantino et al. (2001). Activity determinations were made using 40 ␮l of PMS of the sample, 250 ␮l of NADH (0.24 mM) and 40 ␮l of pyruvate (10 mM) in Tris–NaCl buffer (0.1 M, pH 7.2). ChE activity was determined using acetylthiocholine as substrate and measuring at 414 nm during 5 min the conjugation product between thiocholine (a product of the degradation of acetylthiocholine) and 5,5-dithiobis-2-nitrobenzoic acid (DTNB) (absorbance increase) in phosphate buffer, according to the method of Ellman et al. (1961). Activity determinations were made using 40 ␮l of PMS of the sample, 250 ␮l of reaction mixture (acetylthiocholine (7.5 mM) and DTNB (10 mM)) in phosphate buffer (0.1 M, pH 7.2). GST activity was determined at 340 nm by monitoring the increase in absorbance during 5 min, following the general methodology described by Habig and Jakoby (1981) as modified by Frasco and Guilhermino (2003) Activity determinations were made using 100 ␮l of PMS of the sample and 200 ␮l of reaction mixture (10 mM reduced glutathione (GSH)) and 60 mM 1-chloro-2.4dinitrobenzene in phosphate buffer (0.05 M, pH 6.5). CAT activity was measured at 240 nm by monitoring (every 10 s, for 2 min) the decrease of absorbance due to degradation of H2 O2 , as described by Claiborne (1985). Fifteen microliters of PMS were mixed with 135 ␮l of reaction solution (H2 O2 , 30 mM) and 150 ␮l of phosphate buffer (0.05 M, pH 7.0). Enzymatic activities were determined in quadruplicate and expressed as nanomoles of substrate hydrolyzed per minute (U) per mg of protein. Protein concentration in the samples was determined in quadruplicate by the Bradford method (1976), at 595 nm, using ␥-globulin as standard. All biochemical determinations were made spectrophotometrically in 96 wells microplates using a Thermo Scientific Multiskan® Spectrum. 2.4. Statistical analysis SigmaPlot V.11.0 (SysStat, San Jose, California,USA) statistical package was used for statistical analyses. ANOVA (one-way analysis of variance) was performed followed by a post hoc test to assess differences between treatments and control. Depending if normality and homoscedasticity of data were verified or not a parametric (followed by the post hoc Dunnett’s test) or non-parametric ANOVA (followed by the post hoc Dunn’s test) were chosen respectively. Lethal concentration values (LC50 ) and effect concentration values (EC50 ) were calculated for each parameter by fitting logistic

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Table 1 E(L)C50 values for mortality, hatching and developmental effects observed in zebrafish embryos exposed to 17␣-methyltestosterone. Values were obtained fitting models to data using non-linear regression (± standard error in between brackets). Adjusted model

p(r2 )

Tail malformations 7.83 (0.93) 72 h

Weibull 4-parameter

0.9

Edemas 72 h

Endpoints

Values (mg/l)

7.52 (0.33)

Logistic 4-parameter

0.95

Head abnormal development 9.13 (0.05) 72 h

Weibull 4-parameter

0.98

Hatching delay 72 h 96 h

8.81 (1.76) 8.76 (0.23)

Logistic 4-parameter Logistic 4-parameter

0.89 0.95

Mortality 24 h 48 h 72 h 96 h 120 h

10.42 (0.19) 10.26 (0.16) 10.08 (0.15) 10.09 (0.18) 10.18 (1)

Logistic 4-parameter Logistic 4-parameter Logistic 4-parameter Logistic 4-parameter Logistic 4-parameter

0.98 0.97 0.98 0.97 0.94

p(r2 ) goodness of fit measure, All p(F) test for regression <0.001.

dose-response curves. All statistical analyses were performed with a significance level of 0.05. 3. Results In the present study, fertilized zebrafish eggs were exposed for 120 h to several concentrations of MT. Embryo mortality in the control was below 10% as required for test validity. Moreover organisms from this group presented a normal embryo development as described by Kimmel et al. (1995). LC50 values calculated (Table 1) did not vary significantly along the time of exposure: the 24h-LC50 was 10.42 mg/l while the 120h-LC50 was still 10.18 mg/l. An overview of the mortality and hatching of embryos in the 120 h of the exposure can be seen in Fig. 1. MT caused a dose-dependent hatching delay and/or inhibition. Most of the organisms showing hatching delay did not hatch at all until the end of the test. This was observed for all surviving organisms exposed to 11 mg/l and approximately half of the surviving organisms exposed to 5 mg/l (Fig. 1). Development was affected in embryos exposed to MT. At 24 h a delay in the development of the head was identified (Fig. 2a–c); at 48 h spine deformities, cardiac edemas and eye malformations were recorded in embryos exposed to concentrations above 9 mg/l of MT (Fig. 2d–f). At 72 h the most important effects observed were spine deformity, cardiac edema and malformation of the head at concentrations above 5 mg/l (Figs. 2 g–i and 3 A ). These effects were still observed at 120 h (Fig. 3B). Please refer to Table 1 for EC50 values for development effects. No differences were observed between negative and solvent control for the endpoints evaluated. A sublethal range of concentrations was tested to analyze effects at biochemical level. ChE and LDH activities were be inhibited at the highest concentrations tested while GST and CAT activities were not affected. VTG levels in zebrafish embryos were reduced at all concentrations tested (Fig. 4). Generally, the toxicity of the commercial formulation of MT did not differ from the active compound. In the supplementary material, an overview of the test is shown (Fig. S1). In contrast to the MT, the pattern of response of cMT was time-dependent with LC50 values decreasing from 16.61 mg/l at 24 h to 9.86 mg/l at the end of the test (120 h) (Table S1). Hatching was also inhibited at highest concentrations (Fig. S1) with LC50 values (8.03 and 8.09 mg/l at 72 and 96 h respectively) very similar to what was obtained with the pure compound. Similar patterns were also observed for embryo

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Fig. 1. General overview of 17␣-methyltestosterone effects on zebrafish embryos during the 120 h of exposure. “sol” refers to the solvent control.

4. Discussion

Fig. 2. Malformations observed in embryos exposed to 17␣-methyltestosterone: (a) normal embryos at 24 h, (b and c) delay in head development at 24 h; (d) control at 48 h, (e and f) spine deformity, edema and eye malformation at 48 h; (g) control hatched embryo at 72 h, (h) scoliosis, (i) edema, hatching delay and tail malformation at 72 h.

development and biochemical parameters. Spine deformities, cardiac edemas and abnormal development of the head were the most important developmental effects observed (Fig. S2), while for biochemical parameters reduction in VTG and inhibition of LDH and ChE activities characterized the effects at the highest concentrations. As for MT, no changes in GST or CAT activities were observed in embryos exposed to cMT (Fig. S3).

MT effects are particularly well documented in the development of reproductive structures of non-target organisms. However, effects on the embryonic development of fish species are poorly studied. Boudreau et al. (2005) studied the effects of MT in the embryonic, larval and juvenile stages of the fish Fundulus heteroclitus. After 60 days of exposure, a series of skeletal (scoliosis, malformed fins, facial and jaw abnormalities) and soft tissue (anal swelling) abnormalities were observed in organisms exposed to concentrations above 10 ␮g/l. In zebrafish, there are some evidences of effects in the embryo development elicited by endocrine disruptor compounds although they concern estrogen substances and not androgens like MT (Liu et al., 2010a; Santos et al., 2014). To our knowledge, this work is the first evidence of developmental effects of MT in zebrafish. In invertebrates, however, there are several evidences of the interference of androgens in the embryonic development. For instance, Vogt (2007) exposed eggs of the marbled crayfish to 0.1 mg/l of MT. They observed a decreased hatching success, embryonic development delay, reduced growth, and malformations of the appendages in the juveniles. In contrast to marbled crayfish, zebrafish embryos seem to be more resistant to MT since development effects (including hatching) were only observed at concentrations above 5 mg/l. From the biomarkers evaluated, VTG, LDH and ChE appeared to be sensitive to MT. The presence of VTG in fish at early developmental stages well before the phase of VTG incorporation into female gonads is well documented and has been used as a measure of exposure to endocrine disruptors (Tyler et al., 1999). Early VTG synthesis has been detected through the quantification of VTG expression and/or protein levels as early as in 24 hpf embryos (Lattier et al., 2002; Muncke and Eggen, 2006; Tyler et al., 1999). Our

Fig. 3. Effects (average ± standard error) of 17␣-methyltestosterone on zebrafish embryo development after 72 h (A) and 120 h (B) of exposure. “Sol” refers to the solvent control.

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Fig. 4. Activities of biochemical markers measured in zebrafish embryos after 96 h of exposure to 17␣-methyltestosterone (average ± standard error). Asterisks (*) indicate a result different from the respective control (Dunnet or Dunn’s test; p < 0.005).

findings showed that exposure of zebrafish embryos (0–96 hpf) to MT decreased VTG levels at all the concentrations tested (0.004–5 mg/l MT). Similar results were reported with MT exposure during other stages of fish development such as juveniles and adults. This was the case of juvenile zebrafish exposed to 26–500 ng MT/l from 20 to 60 days post hatching (Örn et al., 2003) and of female medaka O. latipes exposed for 3 weeks to 188 and 380 ng/l MT (Kang et al., 2008). However opposite effects of MT on VTG levels have been also described for zebrafish and other fish species (e.g. Andersen et al., 2006; Hogan et al., 2008). In fact, the ability of MT to affect VTG synthesis has been previously shown to vary with the dose of exposure and to be species-, life stage- and sexdependent (e.g. Andersen et al., 2006; Hornung et al., 2004). For instance, Andersen et al. (2006) obtained increased VTG synthesis in adult male zebrafish with low MT concentrations (4.5 ng/l), but no effect at higher concentrations and have suggested that androgens such as MT may play a role in the regulation of VTG synthesis. The observations that MT can increase VTG levels in zebrafish is consistent with the results obtained for other fish species, namely Cyprinus carpio and fathead minnow (Ankley et al., 2001; Moens et al., 2006; Zerulla et al., 2002). The mechanistic basis of this result might be that MT, apart from the androgenic action, might also induce estrogenic responses when metabolized to the estrogen 17alpha-methylestradiol by aromatase. This estrogen can then lead to the increase of VTG levels (Seki et al., 2004). Since our results are not consistent with an estrogenic mode of action of MT, it can be hypothesized that zebrafish embryos were not able to aromatize MT at the concentrations tested, but this could only be confirmed measuring aromatase activity. The decrease of VTG levels by MT remains unclear and further research should focus on how this androgen acts through estrogen and androgen receptors and regulates VTG levels in early fish life stages. LDH is involved in the anaerobic pathway of energy production, converting pyruvate to lactate in the glycolytic pathway and has been used as a general biomarker of stress in fish (e.g. Vieira et al., 2008). In our study, the highest concentrations of MT tested caused an inhibition of LDH activity. No studies were found where LDH had been analysed in fish early life stages subjected to MT, but Hasheesh et al. (2011) analysed LDH levels in adult tilapia collected from ponds with and without MT treatment detecting no differences between them. However, Peter and Oommen (1989) found an inhibition of liver LDH activity in the fish Anabas testudineus injected with testosterone. These authors suggested that the androgen treatment may result in a shift to the aerobic metabolism as seen by LDH decrease and high rate of mitochondrial respiration. ChEs are a family of enzymes important for neurotransmission, being responsible for the degradation of the neurotransmitter acetylcholine in the cholinergic synapses. MT is not considered a cholinesterase inhibitor (Kegley et al., 2014), however, several works, mainly conducted with rats account for an indirect

hormone-dependent regulation of acethylcholine receptors and cholinesterase isoenzymes (Bleisch et al., 1982; Godinho et al., 1994; Illsley and Lamartiniere, 1981). For instance, in the work of Godinho et al. (1994), after rat castration, the levels of AChE decreased in the androgen-dependent ani muscle; on the other hand, treatment with androgen substances like MT tended to restore AChE activities. In our work, however, a decrease and not an increase in ChE levels was observed after zebrafish embryos exposure to 5 mg/l of MT (16% inhibition of ChE activity when compared to control). No evidences in literature support our result, only one work was found regarding effects of testosterone in ChE of the fish (O. latipes) which reported that muscular ChE of adults was unaffected after treatment with 100 ␮g/l of testosterone for 6 days (El-Alfy and Schlenk, 2002). The inhibition of ChE in this study was observed at very high concentrations of MT (close to lethality levels) and may thus represent an unspecific secondary effect of toxicity. Exposure to MT did not alter the GST and CAT activity. These results suggest that in these exposure conditions, GST did not contribute to the MT excretion in zebrafish early life stages. Furthermore, considering previous studies showing that steroid androgens can contribute to oxidative stress and reactive oxygen species production in rodents (Liu et al., 2010b) and birds (Alonso-Alvarez et al., 2007), changes in CAT activity should be expected in zebrafish embryos. However, since our results did not confirm this response; we hypothesize that the high variability of data prevented the discrimination of effects given that a trend to activity increase can be observed for concentrations above 0.023 mg/l. Higher concentrations should be tested to confirm this trend. In this study, VTG was the most sensitive parameter among the endpoints measured in zebrafish embryos exposed to MT, detecting effects at the lowest concentration tested (4 ␮g/l). This is a very relevant finding since, according to the available literature (Barbosa et al., 2013), 4 ␮g/l is a concentration far below the concentrations that may be detected in water bodies adjacent to aquacultures. Several methods for VTG determination in juvenile and adults are widely applied for endocrine disruption assessment. Response of VTG like-proteins in embryos in our study suggests that, at least in the case of MT, this endpoint can be used to predict the reproductive effects already observed in adult fish with a fastest and most cost effective approach. However, long term studies to follow up the effects observed in embryos until the adult stage would be needed to support this conclusion. Furthermore, in this work a relatively short period of exposure was used, hence, chronic effects might be expected at concentrations far below 4 ␮g/l. Commercial formulations of compounds such as pharmaceuticals or pesticides have very often higher toxicities in non-target organisms than the pure compounds, mainly due to the existence in their composition of other ingredients, generally not taken into account for risk assessment. This seems not to be the case with the formulation tested (cMT) acquired from Bioativa (Brazil), given that

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the toxicity observed was very similar to the toxicity elicited by the pure compound. Overall, our results, especially VTG depletion, indicate that environmental hazards of 17␣-methyltestosterone should be addressed in further studies. More data on biological effects on non-target organisms and data on environmental concentrations are needed to guarantee a sustainable use of MT in aquaculture worldwide. 5. Conclusions Methyltestosterone, both pure and in formulation, were shown in this study to induce several developmental and biochemical effects in zebrafish early life stages. Most important developmental effects included tail malformations, edemas, abnormal development of the head and hatching delay (72 h-EC50 values of 7.83, 7.52, 9.13 and 8.81 mg/l respectively). At biochemical level MT inhibited VTG like-proteins even at the lowest concentration tested (4 ␮g/l) and inhibited cholinesterase and lactate dehydrogenase at the highest concentrations tested (5 mg/l). Although data regarding measured environmental concentrations of MT are scarce, available information indicates that effects in VTG were detected at ecological relevant concentrations, and thus particular importance should be given to this alteration and its further impact on population health. Acknowledgments This work was supported by the Portuguese Foundation for Science and Technology (FCT) through CESAM (UID/AMB/50017/2013) and through the scholarships of ID (SFRH/BPD/90521/2012) and MSM (SFRH/BPD/100448/2014) and the Brazilian Ministry of Science and Technology through the scholarships of RO (CNPq BJT-A, Project No A058/2013), CKG (Bolsista Produtividade CNPq) and AMVMS (Bolsista CAPES/BRASIL, Project No A058/2013). Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.etap.2016.04.014. References Örn, S., Holbech, H., Madsen, T.H., Norrgren, L., Petersen, G.I., 2003. Gonad development and vitellogenin production in zebrafish (Danio rerio) exposed to ethinylestradiol and methyltestosterone. Aquat. Toxicol. 65, 397–411, http:// dx.doi.org/10.1016/S0166-445X(03)00177-2. Alonso-Alvarez, C., Bertrand, S., Faivre, B., Chastel, O., Sorci, G., 2007. Testosterone and oxidative stress: the oxidation handicap hypothesis. Proc. R. Soc. B Biol. Sci. 274, 819–825, http://dx.doi.org/10.1098/rspb.2006.3764. Andersen, L., Goto-Kazeto, R., Trant, J.M., Nash, J.P., Korsgaard, B., Bjerregaard, P., 2006. Short-term exposure to low concentrations of the synthetic androgen methyltestosterone affects vitellogenin and steroid levels in adult male zebrafish (Danio rerio). Aquat. Toxicol. 76, 343–352, http://dx.doi.org/10.1016/ j.aquatox.2005.10.008. Ankley, G.T., Jensen, K.M., Kahl, M.D., Korte, J.J., Makynen, E.A., 2001. Description and evaluation of a short-term reproduction test with the fathead minnow (Pimephales promelas). Environ. Toxicol. Chem. 20, 1276–1290, http://dx.doi. org/10.1002/etc.5620200616. Ankley, G.T., Jensen, K.M., Durhan, E.J., Makynen, E.A., Butterworth, B.C., Kahl, M.D., Villeneuve, D.L., Linnum, A., Gray, L.E., Cardon, M., Wilson, V.S., 2005. Effects of two fungicides with multiple modes of action on reproductive endocrine function in the fathead minnow (Pimephales promelas). Toxicol. Sci. 86, 300–308, http://dx.doi.org/10.1093/toxsci/kfi202. Barbosa, I.R., Lopes, S., Oliveira, R., Domingues, I., Soares, A.M.V.M., Nogueira, A.J.A., 2013. Determination of 17␣-methyltestosterone in freshwater samples of tilapia farming by high performance liquid chromatography. Am. J. Anal. Chem. 04, 207–211, http://dx.doi.org/10.4236/ajac.2013.44026. Bleisch, W.V., Harrelson, A.L., Luine, V.N., 1982. Testosterone increases acetylcholine receptor number in the levator ani muscle of the rat. J. Neurobiol. 13, 153–161, http://dx.doi.org/10.1002/neu.480130207.

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