Effects of nonylphenol and ethinylestradiol on copper redhorse (Moxostoma hubbsi), an endangered species.

Ecotoxicology and Environmental Safety 108 (2014) 168–178

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Effects of nonylphenol and ethinylestradiol on copper redhorse (Moxostoma hubbsi), an endangered species Domynick Maltais n, Robert L. Roy Pêches et Océans Canada, Institut Maurice-Lamontagne, 850 route de la Mer, Mont-Joli, QC, Canada G5H 3Z4

art ic l e i nf o

a b s t r a c t

Article history: Received 26 March 2014 Received in revised form 30 June 2014 Accepted 3 July 2014

The copper redhorse, Moxostoma hubbsi, is an endangered species endemic to Quebec. The presence of contaminants, in particular endocrine disrupting chemicals (EDCs), in its habitat has been advanced as partly responsible for the reproductive difficulties encountered by the species. In the present study, immature copper redhorse were exposed to the estrogenic surfactant nonylphenol (NP; 1, 10 and 50 mg/l) and the synthetic estrogen 17α-ethinylestradiol (EE2; 10 ng/l) for 21 days in a flow-through system. The endpoints investigated included general health indicators (hepatosomatic index and hematocrit), thyroid hormones, sex steroids, brain aromatase activity, plasma and mucus vitellogenin (VTG), cytochrome P4501A protein expression and ethoxyresorufin-O-deethylase activity, heat shock protein 70 (HSP70) and muscle acetylcholinesterase. Exposure to 10 ng EE2/l significantly increased brain aromatase activity. Exposure to 50 mg NP/l resulted in a significant reduction of plasma testosterone concentrations and a significant induction of hepatic HSP70 protein expression. NP at 50 mg/l also induced plasma and mucus VTG. The presence of elevated VTG levels in the surface mucus of immature copper redhorse exposed to NP, and its correlation to plasma VTG, supports the use of mucus VTG as a non-invasive biomarker to evaluate copper redhorse exposure to EDCs in the environment and contribute to restoration efforts of the species. The results of the present study indicate that exposure to high environmentally relevant concentrations of NP and EE2 can affect molecular endpoints related to reproduction in the copper redhorse. Crown Copyright & 2014 Published by Elsevier Inc. All rights reserved.

Keywords: Copper redhorse Endocrine disrupting chemicals Vitellogenin Aromatase Heat shock protein 70 Steroid hormones

1. Introduction The copper redhorse, Moxostoma hubbsi, is a catostomid fish (cypriniformes order). It is endemic to Quebec with a distribution range restricted to a section of the St-Lawrence River, between Lake Saint-Louis and Lake Saint-Pierre, and some of its tributaries (DFO, 2011). Genetic analysis, telemetric monitoring, and studies conducted on contamination levels present in the copper redhorse indicate that the species now forms one single population (de Lafontaine et al., 2002; Lippé et al., 2006; Gariépy, 2008). The copper redhorse has a long lifespan (over 30 years) and a specialized diet, feeding almost exclusively on small molluscs (DFO, 2011). The size of the spawners is generally greater than 500 mm and spawning occurs between midJune and early July (Mongeau et al., 1992). This species is a nonguarding, open substrate and lithophil spawner (Coker et al., 2001). The copper redhorse has been listed as endangered under the Committee on the Status of Endangered Wildlife in Canada (COSEWIC, 2004), as vulnerable under the IUCN (International Union for

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Corresponding author. Fax: þ1 418 775 0718. E-mail addresses: [email protected], [email protected] (D. Maltais). http://dx.doi.org/10.1016/j.ecoenv.2014.07.004 0147-6513/Crown Copyright & 2014 Published by Elsevier Inc. All rights reserved.

Conservation of Nature) Red List of Threatened Species and threatened under the Quebec Act respecting threatened or vulnerable species since 1999. The population has difficulties in reproducing naturally, a problem that has been tentatively linked to the presence of contaminants, in particular endocrine disrupting chemicals (EDCs), in the habitat of the species (Gendron and Branchaud, 1997; de Lafontaine et al., 2002). Contaminants have been singled out as a threat to the recovery of the copper redhorse (COSEWIC, 2004). A priority of the past and current provincial recovery plans and the federal recovery strategy for the copper redhorse is to evaluate exposure to EDCs and the possible role of EDCs in the reproductive problems encountered by the species (Équipe de rétablissement du chevalier cuivré, 2004, 2012; DFO, 2012). Gendron and Branchaud (1997) hypothesized that metabolites of alkylphenol polyethoxylates (APEs) might impair final gamete maturation in the copper redhorse. Moreover, a study on spottail shriners (Notropis hudsonius) suggested that sexual differentiation and reproductive functions of fish may be altered in the habitat of the copper redhorse (Aravindakshan et al., 2004). APEs include nonylphenol ethoxylates (NPEs), non-ionic surfactants widely used in products for agricultural, industrial, commercial and domestic applications. The production of NPEs involves the use of nonylphenol (NP). NP is introduced in the aquatic environment during its use and as a degradation product of NPEs (Vazquez-Duhalt et al.,

D. Maltais, R.L. Roy / Ecotoxicology and Environmental Safety 108 (2014) 168–178

2005; Soares et al., 2008). NP is found in freshwaters, estuaries and sediments around the world, with concentrations between 0.7 and 15 mg/l in river water (Soares et al., 2008). While NP concentrations o1 mg/l are frequent in surface waters (Soares et al., 2008), higher concentrations have been measured in Spain (600 mg/l, Solé et al., 2000a) and US waters (95 mg/l, Dachs et al., 1999). In the Richelieu river, an important habitat of the copper redhorse and the only watercourse where reproductive activity has been reported, Berryman et al. (2004) detected a yearly average level of NPEs of 1.66 μg/l, 0.35 mg/l in NP equivalent, (range of 0.62–3.71 μg/l; 0.029– 0.74 mg/l in NP equivalent) at Otterburn Park. The presence of nonylphenolic compounds was also reported in St. Lawrence River surface water (up to 9.56 mg/l; 2.06 mg/l in NP equivalent) and sediments downstream of the Montreal Urban Community discharge point (Berryman et al., 2004; Bennie et al., 1997). APEs were also found in the sediments of the St. Lawrence River at several sites around the Island of Montreal (Sabik et al., 2003). High concentrations of nonylphenolic compounds (up to 55.3 mg/l; 8.62 mg/l in NP equivalent) were measured in the Yamaska River (Berryman et al., 2004), a tributary of the St. Lawrence River that was part of the historical habitat of the copper redhorse. However, its presence is now uncertain and even unlikely (DFO, 2011). NP has a structural similarity with 17β-estradiol (E2; Soares et al., 2008). Studies in vivo and in vitro involving exposures to NP have shown that it acts as an estrogen mimic in fish, with effects mediated by the estrogen receptor (ER) (Vazquez-Duhalt et al., 2005; Soares et al., 2008). Moreover, Lee et al. (2003) showed that NP has an antiandrogenic activity at multiple steps of androgen receptor activation and function. Alteration of endocrine function can have effects on critical reproductive stages like sexual differentiation/puberty, gamete production and gonadal growth. The effects of EDCs on fish reproduction can be investigated by measurement of reproductive biomarkers. The induction of the female egg yolk precursor vitellogenin (VTG) in male or immature fish is an established biomarker of exposure to estrogenic compounds in aquatic environment and one of the most commonly investigated endpoint in response to NP exposure (Coady et al., 2010). Alterations of steroid hormone levels in fish blood have also been reported following NP exposure (Kortner et al., 2009; Sayed et al., 2012) and linked with decreased fecundity among fish populations (Coady et al., 2010). The cytochrome P450 aromatase (CYP19) is an important steriodogenic enzyme catalyzing the conversion of androgens, like testosterone, to estrogens, like estradiol (Diotel et al., 2010). Aromatase can be found in various organs including gonads and brains and has a crucial role in development, sexual differentiation, reproduction and behavior (Diotel et al., 2010). Unlike mammals, teleost fish exhibit an intense aromatase activity in brain (Diotel et al., 2010). Alterations of brain aromatase activity were observed in Atlantic salmon (Salmo salar) following NP waterborne exposures (Kortner et al., 2009). The cytochrome P4501A (CYP1A) is an enzyme responsible for the metabolism of xenobiotics. Its ethoxyresorufin-O-deethylase (EROD) activity is used as a biomarker of exposure to aryl hydrocarbon receptor (AhR) ligands in aquatic environments (Goksoyr and Forlin, 1992). Studies have suggested an interaction between the systems responsible for EROD induction and VTG production in fish (Arukwe et al., 1997; Kirby et al., 2007). Estrogenic substances, including NP, has been shown to affect the synthesis of CYP1A and EROD activity in diverse fish, presumably due to a crosstalk between the ER and AhR (Arukwe et al., 1997; Solé et al., 2000b; Cionna et al., 2006; Sturve et al., 2006; Kirby et al., 2007). Compared to its estrogenic effects, little is known about potential other effects of NP on aquatic animals. Data suggest that NP can interfere with thyroid hormones functions (Carlisle et al., 2009). Thyroid hormones are important in fish physiology (metabolism, growth, reproduction and brain development) and have been

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proposed as potential early biomarkers of NP endocrine disruption (Zaccaroni et al., 2009). Heat shock proteins (HSPs) are crucial for protein assembly, maintenance of protein integrity, protein translocation and processing of misfolded proteins due to various kinds of stress (Iwama et al., 1998). These proteins also play an important role in steroid-receptor interaction and function (Iwama et al., 1998). HSPs, particularly HSP70, have been shown to be expressed in fish in response to toxicants, including EDCs, and proposed as biomarkers for pollutant exposure (Vijayan et al., 1998; Iwama et al., 1998; Aït-Aïssa et al., 2000; Carnevali and Maradonna, 2003). Acetylcholinesterase (AChE) is an enzyme that catalyzes the hydrolysis of the neurotransmitter acetylcholine. ChE activity has been shown to be affected by ionic surfactants (Guilhermino et al., 2000; Stock et al., 2004). NP was reported to inhibit AChE activity of rat pheochromocytoma PC12 cells (Talorete et al., 2001). Moreover, ChE inhibition was observed in muscle of guppies (Poecilia reticulata) exposed to NP (Li, 2008). 17-α-Ethinylestradiol (EE2) is a synthetic estrogen used in oral contraceptives. EE2 is only partially metabolized and not fully removed by wastewater treatment and, consequently, is detected at low ng/ml concentrations in discharges from treatment plants (Ternes et al., 1999; Baronti et al., 2000) and in river waters (Cargouët et al., 2004; Williams et al., 2003; Zuo et al., 2006). The presence of EE2 in the effluents of the Montreal Urban Community treatment plant discharged in the St. Lawrence River has been reported (Aravindakshan et al., 2004). EE2, at environmentally relevant concentrations, is known to induce VTG, brain aromatase activity and intersex gonads in male fish (Hallgren and Olsen, 2010; Christen et al., 2010). The purpose of this study was to provide an overview of the effects of exposure to environmentally relevant concentrations of NP and EE2 on copper redhorse physiology. Given the precarious status of this species, it is not possible to perform terminal toxicity bioassays on copper redhorse captured in the wild. This study takes advantage of a unique opportunity to study the effects of endocrine disrupting compounds on a limited number of copper redhorse raised in captivity and donated by the Montreal Biodome. The endpoints included general health indicators (hepatosomatic index and hematocrit), thyroid hormones (triiodothyronine, thyroxine), sex hormones (estradiol, testosterone and 11-ketotestosterone), brain aromatase activity, VTG in plasma and surface mucus, CYP1A, EROD, HSP70 and muscle acetylcholinesterase (AChE). This study provides new information about the effects of environmental estrogens on a redhorse species and contributes to investigate the role of EDCs in the difficulties facing copper redhorse reproduction. In addition, this work validates a noninvasive tool, the measurement of VTG in mucus, for monitoring exposure to EDCs in this endangered species.

2. Materials and methods Except where mentioned, all chemicals were obtained from Sigma-Aldrich (Oakville, ON, Canada). The stock solution of NP (technical grade; CAS number 84852-15-3) and EE2 were prepared in 100 percent ethanol to achieve nominal concentration in the tank and 0.004 percent ethanol. Protein concentrations in samples and homogenates were determined by the Bradford method using bovine serum albumin as standard (Bio-Rad laboratories, Mississauga, ON, Canada).

2.1. Fish A limited number of immature copper redhorse, raised in captivity, were provided by the Montréal Biodome (Montreal, QC, Canada). These unique fish were F1 offspring of wild caught copper redhorse. They had scoliosis and were not intended to be released in the wild. Despite of their scoliosis, fish were in good condition and able to eat properly. At the Institut Maurice-Lamontagne (IML), fish were maintained in tanks supplied with flowing de-chlorinated municipal tap water at controlled temperature (12–18 1C depending on the season). Fish were fed sinking feed (Martin Mills, Elmira, ON, Canada) four to five times a week.

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2.2. Exposure and experimental set-up All practices (maintenance and experimental manipulations) of the animals were approved by the Animal Care Committee of the IML, according to guidelines of the CCAC (Canadian Council on Animal Care, 2005). Five days prior to the exposures, 34 copper redhorse (27.070.4 cm length; 337.6717.9 g weight; mean7S.E.M.) were anesthetized, measured, weighed and sampled for blood (0.25 ml/kg body weight) and surface mucus, as described below. The fish were randomly transferred and acclimated for 5 days in six 120 l Swedish tanks (5–6 fish per tank) supplied with 2.5 l/min flowing de-chlorinated municipal tap water at 13 1C, with a photoperiod of 16:8 h (light:dark). Water and photoperiod parameters were similar to those before transfer. After acclimation, fish were exposed for 21 days to NP at nominal concentrations of 1, 10, 50 mg/l, EE2 at 10 ng/ml, the carrier solvent (0.004 percent ethanol) and a water control. Concentrated solutions of NP (25, 250 and 1250 mg/l) and EE2 (250 mg/l) were delivered by peristaltic pumps (Econopump, BioRad laboratories, Mississauga, ON, Canada) at a flow rate of 0.1 ml/min and mixed online with de-chlorinated municipal tap water (2.5 l/min) to the target concentrations. Physico-chemical characteristics of the water were monitored daily and temperature and dissolved oxygen (means7S.D.) were maintained at 13.270.1 1C and 9.4670.40 mg/l, respectively. Fish were fed daily ad libitum as recommended in the OECD (Organisation for Economic Co-operation and Development) guideline for the testing of chemicals (OECD, 2009). No aggressive, territorial or competitive behaviors were observed during the experiment. Uneaten food and waste were removed every day. Fish were not fed the day prior to sampling/necropsy to prevent vomiting and fecal contamination in the anesthesia tank (Canadian Council on Animal Care, 2004; OECD, 2009).

2.3. Sample collection Samples of plasma and surface mucus were collected 5 days before the beginning of the experiment and at the end of the experiment, on day 21. Fish were anaesthetized by immersion in a tricaine methane sulfonate solution (MS 222, 75 mg/l, pH adjusted to 7.0 with 10 percent (w/v) Na2CO3). Surface mucus was collected as described by Maltais and Roy (2007). Blood was collected from the caudal vein using heparinized syringes, as described by Maltais et al. (2010). Blood samples were transferred into tubes containing heparin and aprotinin and centrifuged (3500g, 4 1C, 10 min) to obtain plasma. Plasma samples were stored at 80 1C until analyzed. On day 21, prior to centrifugation, blood samples were also collected, in triplicate, in heparinized microhematocrit tubes (BD, Mississauga, ON, Canada) and centrifuged to determine hematocrits. On day 21, liver, gonads, brain and muscle samples were removed from each fish. Liver and gonad were weighed to calculate hepatosomatic (HSI) and gonadosomatic (GSI) indices by using following formula: weight of gonad or liver 100/weight of fish. Liver, brain and muscle samples were stored at 80 1C until analyzed. 2.4. Water concentrations of NP and EE2 Water samples were collected at least twice a week to confirm actual concentrations of NP and EE2 in exposure water. Water samples were filtered through 1 mm glass-filters and frozen at 20 1C in chemical-free glass bottles until analyzed by ELISA. Except for the 50 mg NP/l samples, thawed water samples were extracted and concentrated on SPE cartridges before analyses. Spiked control water (1 and 10 mg NP/l; 10 ng EE2/l) was used to determine the recovery rate of the extraction/concentration procedures and to correct values. Water samples intended for NP measurements were extracted on Nexus cartridges (6 ml, 200 mg, Varian Canada Inc, Mississauga, ON, Canada), preconditioned with dichloromethane (10 ml), methanol (5 ml) and nanopure water (5 ml). Samples were loaded at a flow rate of approximately 10 ml/min. Cartridges were washed with nanopure water (5 ml) and 50 percent methanol solution (5 ml) and then dried under vacuum for at least 45 min. The analyte was eluted with dichloromethane (6 ml) and recuperated in glass tubes. The solvent was evaporated to dryness under a stream of nitrogen and the residue was dissolved with DMSO and methanol at a volume ratio of 1:10. The mixture was stirred on a multitube vortexer and water was added to adjust the content to 1 percent DMSO and 10 percent methanol. NP was measured using an alkylphenol (AP) ELISA kit (Japan EnviroChemicals, Ltd. Tokyo, Japan). The limit of quantification (LOQ) of the assay was 5 mg/l. The recovery percentage of the extraction/ concentration procedure was 67.173.7 (mean7S.D.; n¼ 3) for the 1 mg/l spiked control water and 84.776.3 (mean7S.D.; n¼ 3) for the 10 mg/l spiked control water. In nonextracted samples (50 mg NP/l), DMSO and methanol were added to samples to final concentrations of 1 percent and 10 percent (v/v), respectively, and the mixture was assayed. Water samples intended for EE2 measurements were extracted on C18 SPE columns (6 ml, 500 mg; Mallinckrodt Baker, Phillipsburg, NJ, U.S.A) preconditioned with methanol (5 ml) and water (10 ml). After passing samples through the columns, the cartridges were washed with nanopure water (5 ml), dried by maintaining suctioning for 5 min, and further washed with hexane (5 ml). The analyte was eluted with dichloromethane (5 ml) and recovered in glass tubes. The solvent was evaporated to dryness under a stream of nitrogen and the residue was dissolved in 100 percent methanol and stirred on a multitube vortexer. Nanopure water was added to adjust the content to 10 percent

methanol. EE2 was measured using an EE2 ELISA kit (Japan EnviroChemicals, Ltd. Tokyo, Japan). The LOQ of the assay was 0.05 mg/l. The recovery percentage of the extraction/ concentration procedure was 88.479.1 (mean7S.D.; n¼11). 2.5. Thyroid hormones analysis Plasma collected on day 21 of the experiment was assayed for total triiodothyronine (T3) and thyroxine (T4), using commercially available ELISA kits (Monobind, Lake Forest, CA, U.S.A.). 2.6. Steroid analyses Plasma concentrations of E2, T and 11-KT were measured with enzyme immunoassays (EIAs, Cayman Chemical, Ann Arbor, Michigan, USA). Frozen plasma samples were thawed on ice. 11-KT samples were diluted 1:10 in EIA buffer, heated at 80 1C for 30 min and then centrifuged 10 min at 10,000g. The supernatant was recuperated and assayed for 11-KT. Samples for E2 and T determination (100 ml) were spiked with tritiated-E2 and tritiated-T, respectively (100 ml, 2000 cpm/100 ml; GE healthcare, Montreal, Canada). Nanopure water (300 ml) was added, and the solution was vortexed and incubated 30 min at room temperature. Ether (2.5 ml) was added and the tubes were vortexed 30 min on a multitube vortexer. The aqueous phase was frozen in liquid nitrogen whereas the organic phase was transferred to a glass tube coated with Sigmacote (Sigma Aldrich, St-Louis, MO, USA). This step was repeated two times. The organic phase was then evaporated under a stream of nitrogen at room temperature. The residue was dissolved in EIA buffer (500 ml) and assayed by EIA. 50 ml of the solution was mixed with scintillation cocktail and counted on a Beckman LS6000TA counter (Beckman Coulter Canada, Montreal, QC, Canada). Extraction efficiency was determined for each sample using % recovery of radiolabelled spikes. E2 and T concentrations obtained by EIA were then corrected for the extraction recovery. 2.7. Vitellogenin analysis Plasma and surface mucus VTG concentrations were measured using a commercially available carp VTG ELISA (Biosense Laboratories, Bergen, Norway) and purified copper redhorse VTG standard as previously described by Maltais et al. (2010). VTG concentrations in surface mucus were normalized to total protein concentration to eliminate possible bias caused by variations in the amount of mucus and water at the surface of fish. 2.8. Brain aromatase activity Aromatase activity was quantified using the tritiated water releasing method adapted from Gonzalez and Piferrer (2002), Patel et al. (2006), Villeneuve et al. (2006) and Gonçalves et al. (2008). Tritiated water and [1β-3H]-androstenedione (henceforth 3H-androstenedione) were purchased from Perkin Elmer (Woodbridge, ON, Canada). Brains from individual fish were homogenized at 4 1C in 10 mM potassium phosphate buffer (K2HPO4, pH 7.4, with 100 mM KCl, 1 mM EDTA and 1 mM dithiothreitol), in a proportion of 10 percent (m/v). The homogenates were centrifuged (10 min at 4 1C, 10,000g) to remove insoluble material. A 100 ml volume of the supernatants was mixed with 300 ml of the reaction medium (10 mM K2HPO4, 100 mM KCl, 1 mM EDTA, 1 mM dithiothreitol, pH 7.4, 1 mM NADPH, 10 mM glucose6-phosphate, 1 U/ml glucose-6-phosphate dehydrogenase, 150 nM 3H-androstenedione) in borosilicate tubes and incubated at 28 1C for 90 min in a water bath with shaking. All samples were assayed in duplicates with a blank containing buffer instead of sample, an extraction control with tritiated water instead of the reaction medium and a specificity control containing a control sample and reaction buffer with 100 mM formestane, an aromatase inhibitor. The tubes were immersed in ice-cold water and 200 ml of 30 percent trichloroacetic acid was added. The mixtures were vortexed, incubated at room temperature for 5 min and centrifuged at room temperature (10 min, 2000g); 555 ml of the supernatants was recuperated in glass tubes and extracted with 2 ml chloroform. After vortexing for 60 s, the tubes were centrifuged (10 min, 2000g); 400 ml of aqueous phases was mixed with 400 ml of 5 percent charcoal/0.5 percent dextran in water, incubated 15 min at room temperature and centrifuged (15 min, 2000g). The radiochemical content remaining in the aqueous fraction (tritiated water) was quantified by liquid scintillation counting (400 ml mixed with 5 ml liquid scintillation cocktail). Aromatase activity, expressed as fmol/h/mg protein, was determined by calculating the amount of tritiated water produced and by correcting for the tritium estimated in the blank tubes, the dilution factors, the 3H recovery rate and the protein concentration. Formestane completely inhibited the triated water releasing activity, confirming the specificity of the assay for aromatase activity. 2.9. Western blot analysis of HSP70 and CYP1A Liver samples were homogenized and the cytosolic and the endoplasmic reticulum fractions were prepared using the Endoplasmic Reticulum Isolation Kit from SigmaAldrich. Livers were homogenized in an isotonic extraction buffer (10 mM HEPES, pH 7.8, 250 mM sucrose, 25 mM potassium chloride and 1 mM EGTA; 3.5 ml per g of tissue)

D. Maltais, R.L. Roy / Ecotoxicology and Environmental Safety 108 (2014) 168–178 using a glass homogenizer and an overhead motor. The homogenates were centrifuged at 1000g for 10 min at 4 1C. The supernatants were collected and centrifuged at 12,000g for 15 min at 4 1C. The supernatants were transferred to other tubes and centrifuged at 100,000g for 60 min at 4 1C (Optima MAX-XP ultracentrifuge, Beckman). The supernatants (cytosolic fractions) were collected, aliquoted and stored at 80 1C until HSP70 analysis. The pellets (crude microsomal fractions) were suspended in isotonic extraction buffer (0.3 ml per g of original tissue), aliquoted and stored at 80 1C until CYP1A analysis. Except where indicated, all materials and reagents for electrophoresis and Western blots were supplied by Invitrogen Canada (Burlington, ON, Canada). Samples of liver cytosol and microsomal fraction were thawed, and the protein concentration in each of them was determined. The samples were prepared in NuPAGEs lithium dodecyl sulfate sample buffer containing NuPAGE sample reducing agent (40 mg protein per 15 ml of preparation). Samples were heated to 70 1C for 10 min and loaded into wells (20 mg protein for cytosolic fractions and 13.3 mg protein for microsomal fractions). Sodium dodecyl sulphate polyacrylamide gel electrophoresis (SDS-PAGE) was conducted with NuPAGE Novex 4–12 percent Bis-Tris gels and NuPAGE 4-Morpholinepropanesulfonic acid (MOPS) SDS running buffer containing NuPAGE Antioxidant. Gels were run in a Novex X-Cell Surelock Mini-Cell apparatus at 200 V. Relative MWs were estimated by running the Novexs Sharp Pre-stained Protein Standard. After SDS-PAGE, proteins were electroblotted from the gel onto a polyvinylidene fluoride (PVDF) membrane (Immobilon-P, Millipore) using an XCell II™ Blot Module. The transfer was done at 30 V for 1 h using NuPAGE transfer buffer with 10 percent methanol and 0.1 percent NuPAGE Antioxidant. Detection was achieved with a protein detector TMB™ Western blot kit (KPL, Inc., Gaithersburg, MD, USA). An anti-HSP70 rabbit polyclonal antibody (Cayman Chemical; product code: 19015) diluted 1:2000 was used for HSP70 detection in cytosolic fractions. An anti-CYP1A (fish) rabbit polyclonal antibody (CP-226; Cayman Chemical; product code: 173135) diluted 1:250 was used for CYP1A detection in microsomal fractions. A goat anti-rabbit IgG conjugated to peroxidase, supplied with the kit, was used as secondary antibody. Blots were incubated with TMB (3,30 ,5,50 -tetramethylbenzidine) peroxidase substrate solution. Once developed, the membranes were rinsed in ultrapure water to stop the reaction and dried. The intensity of the HSP70 or CYP1A bands on scanned membranes was measured by densitometry using Quantity Ones 1-D analysis software from Bio-Rad Laboratories. HSP70 or CYP1A levels are expressed as relative intensity/mg protein. The intensities of the bands on blots were normalized with a reference sample (a sample from a control group). The reference sample for HSP70 or CYP1A was included on every gel as an internal standard to correct variability between blots. 2.10. EROD activity in liver EROD activity in the liver microsome fractions was measured at 22 1C by a microplate spectrofluorometric assay using a Fluoroskan Ascent microplate fluorometer (Thermo Scientific, Waltham, MA, USA) with a 530 nm excitation filter and a 585 nm emission filter (adapted from Fragoso et al., 1998). Reaction mixtures contained 50 ml of microsome fraction diluted in 4-(2-hydroxyethyl)-1-piperazineethanesulfonic acid buffer (0.1 M, pH 7.8), 10 ml nicotinamide adenine dinucleotide (20 mg/ml) and 50 ml of 7ethoxyresorufin (0.024 mg/ml in dimethylsulfoxyde). Activity was quantified by measuring the fluorescence of resorufin at 1 min intervals over a total scan time of 13 min. The readings of fluorescence were compared to a resorufin standard curve. All samples were assayed in triplicate. EROD activity was calculated as pmol resorufin/mg protein/ min. 2.11. AChE activity Muscle samples were homogenized 1:10 (w:v) in 50 mM sodium phosphate buffer (pH 7.5) containing 0.1 percent TritonX-100. The homogenates were centrifuged at 10,000g for 20 min at 4 1C and the supernatants (S10 fractions) were recuperated and stored at 80 1C until AChE analysis. AChE activity was measured in 96 well microplates using the colorimetric method of Ellman et al. (1961). The S10 fractions (40 ml) were preincubated with 20 ml of 10 mM tetraisopropyl pyrophosphoramide (isoOMPA), a butyrylcholinesterase inhibitor, under shaking for 25 min at room temperature. Two hundred ml of 50 mM sodium phosphate buffer (pH 7.5) and 10 ml of dithiobisnitrobenzoate (DTNB) were added and the mixtures were incubated under shaking for 5 min at room temperature. The reaction was started by the addition of 30 ml of substrate (acetylthiocholine iodide, 12 mM). The change in absorbance at 405 nm was followed for 5 min. The AChE activity was expressed as nmol/min/mg protein. No butyrylcholinesterase activity was detected in the samples. Replacement of iso-OMPA with 0.1 mM eserine, an AChE inhibitor, completely inhibited (495 percent) the cleavage of acetylthiocholine, confirming that the ChE activity measured in our samples referred to AChE activity. 2.12. Statistical analysis Data were statistically analyzed with Addinsoft XLSTAT version 2011.05.01. Data was tested for normality (Shapiro–Wilk’s test) and homogeneity of variances (Bartlett’s test). When the assumptions were met, a one-way analysis of variance

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(ANOVA), followed by Dunnett’s multiple comparison test, was performed to compare the groups. Otherwise, the nonparametric Kruskal–Wallis test was performed followed by pairwise comparisons using Dunn test with Bonferroni adjustment. Correlation between variables was assessed by Kendall’s tau coefficient. Statistical differences were considered to be significant at p o0.05.

3. Results 3.1. Chemical analyses Water samples collected during the experiment were analyzed for NP and EE2. Mean concentrations (7S.D.) of NP in tanks were 0.85 70.04 mg/l (85 percent of nominal, n¼ 6), 10.58 70.72 (106 percent of nominal, n ¼6) and 47.86 76.34 mg/l (96percent of nominal, n ¼7), respectively. Mean concentration ( 7S.D.) of EE2 in the tank was 9.40 7 1.21 ng/l (94 percent of nominal, n ¼ 8). No NP and EE2 was detected in the water and solvent control. Nominal NP and EE2 values are used in the following text. 3.2. Mortality, somatic indices and hematocrit There was no fish mortality during the exposure period. No significant differences in HSI or GSI were observed between controls and exposed fish (Table 1). Fish used in this study were immature and had very small or non-detectable gonads. None of the treatment significantly influenced hematocrit (Table 1). 3.3. Concentrations of thyroid hormones in plasma Total T3 and T4 were measured in plasma samples collected at the end of the exposure period. Total T4 could not be quantified in all plasma samples. More than half of the plasma samples showed T4 concentrations below the detection limit of the ELISA kit (o0.04 mg/dl) (data not shown). NP and EE2 did not affect T4 concentrations, as hormone concentrations in exposed groups were statistically similar to those measured in fish from the water and solvent controls (Kruskal–Wallis, p 40.05). All samples analyzed had total T3 levels greater than the limit of quantification. T3 average concentrations for the different group ranged between 1 and 2.5 ng/ml. T3 concentrations were not affected by NP or EE2 exposure (data not shown, Kruskal–Wallis, p4 0.05). 3.4. Concentrations of sex steroids in plasma After 21 days of exposure, concentrations of E2 in plasma were less than the detection limit of 35 pg/ml in all fish. T concentrations in fish exposed to 50 mg NP/l were significantly lower than concentrations found in fish from the water and solvent control groups (Table 2). Fish exposed to NP tented to have lower concentrations of 11-KT compared to the controls, but the differences were not significant (Table 2). There were no significant differences in plasma T and 11-KT concentrations between the fish exposed to 10 ng/l EE2 and those in the controls (Table 2). 3.5. Aromatase activity There were no significant differences in brain aromatase activity between fish exposed to NP and those in control groups (Table 2). However, exposure to 10 ng/l EE2 significantly increased brain aromatase activity (Table 2). 3.6. Vitellogenin concentrations in plasma and surface mucus VTG concentrations were measured in plasma and surface mucus extract before and after 21 days of exposure to NP and EE2 (Fig. 1). Concentrations of vitellogenin in plasma and mucus

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five days before beginning the experiment were statistically similar among treatments. Exposure to 50 mg NP/l for 21 days produced a significant increase ( 410 fold) of VTG in plasma as well as in surface mucus (Fig. 1). Plasma and mucus VTG concentrations showed no significant difference between controls and treatments other than 50 mg NP/l. VTG levels in plasma and mucus extract collected at the end of exposure were strongly correlated (Fig. 2; Kendall’s tau, p o0.0001). A strong correlation was also observed between VTG levels in plasma and mucus extract collected before the exposure (data not shown; Kendall’s tau; p o0.0001). 3.7. CYP1A protein expression and EROD activity in liver Copper redhorse CYP1A appeared as a single band of 50 kDa on Western blots (Fig. 3A). CYP1A protein expression in the microsomal fraction of fish exposed to NP or EE2 was not significantly different from controls (Fig. 3A). EROD activity in the microsomal fraction was also not significantly different from controls (Fig. 3B). CYP1A protein levels strongly correlated with EROD activity (Kendall’s tau, p o0.0001). 3.8. HSP70 HSP70 from copper redhorse appeared as a single band of approximately 70 kDa on Western blots (Fig. 4). The quantitative determination of HSP70 bands showed an induction of HSP70 protein by NP. Exposure to 50 mg NP/l for 21 days significantly enhanced (1.5 fold) the expression of this protein (Fig. 4). HSP70 expression was not significantly affected by EE2 (Fig. 4). 3.9. Muscle AChE activity AChE activity was similar for all treatment groups (Kruskal–Wallis, p40.05). In fish exposed to the dilution water, muscle AChE activity (mean7S.E.M.) was 160772 nmol/min/mg, not different than levels in fish exposed to the solvent control (136756 nmol/min/mg). Muscle AChE activities in fish exposed to 1, 10, 50 mg NP/l were 142758, 160765 and 146760 nmol/min/mg, respectively. Finally, in copper redhorse exposed to 10 ng EE2/l, muscle AChE activity was 165774 nmol/min/mg.

4. Discussion In the present study, we evaluated the effects of NP and EE2 at environmentally relevant concentrations on biomarker responses in the endangered copper redhorse. The 96-h LC50 of NP for fish ranges from 17 to 3000 mg/l (reviewed in Servos, 1999). No mortality was observed during the course of our experiment indicating that the concentrations of NP tested were below lethal concentrations for acute toxicity to immature copper Table 1 Hepatosomatic index (HSI), gonadosomatic index (GSI) and hematocrit in immature copper redhorse exposed to NP and EE2 for 21 days. Data are expressed as means 7 S.E.M. (n¼ 5–6). Treatment Water control Solvent control 1 mg NP/l 10 mg NP/l 50 mg NP/l 10 ng EE2/l

HSI (%)

GSI (%)

1.08 7 0.09 0.99 7 0.11 0.99 7 0.10 1.337 0.15 1.317 0.10 1.067 0.12

0.217 0.10 0.377 0.15 0.22 7 0.09 0.26 7 0.11 0.40 7 0.16 0.39 7 0.17

Hematocrit (%) 30.0 7 1.9 30.17 1.0 32.4 7 1.7 27.3 7 2.6 30.9 7 2.5 31.7 7 1.7

There were no significant differences between exposed groups and controls (HSI and hematocrit: ANOVA, p 40.05; GSI: Kruskal–Wallis, p4 0.05).

redhorse. The EE2 concentration tested in this study was several orders of magnitude below the median 96-h LC50 reported for zebrafish, Danio rerio (1.7 mg/l, Versonnen et al., 2003). Waterborne exposure to NP did not affect HSI in copper redhorse, as reported previously in a number of other fish species (Villeneuve et al., 2002; Li and Wang, 2005; Schoenfuss et al., 2008; Zha et al., 2007). An elevated HSI was observed in flounder, Platchthys flesus, exposed to 10 ng EE2/l (Allen et al., 1999). Unlike flounder, copper redhorse exposed to 10 ng EE2/l had HSI similar to control groups, as reported for three-spined stickleback (Gasterosteus aculeatus, Andersson et al., 2007), Chalcalburnus tarichi (Kaptaner et al., 2009) and rainbow trout (Schultz et al., 2003) exposed to similar or higher EE2 concentrations. Some studies have reported alteration of GSI in fish exposed to NP (Jobling et al., 1996; Harries et al., 2000; Zha et al., 2007; Sayed et al., 2012) or EE2 (Scholz and Gutzeit, 2000; Zha et al., 2007). However, other studies have failed to demonstrate a relation between GSI and NP exposure (Villeneuve et al., 2002; Schoenfuss et al., 2008), and Andersson et al. (2007) observed no effects on GSI in female threespined stickleback exposed to 50 ng EE2/l for 21 days. In the present study, NP and EE2 had no effect on copper redhorse GSI, perhaps due to the immature status of the fish. The physiology, health and activity can affect hematocrit levels of fish. In this study, copper redhorse hematocrits were not altered by NP exposure, as reported in carp, Cyprinus carpio, exposed for 70 days to 1 to 15 mg NP/l or injected with EE2 (Schwaiger et al., 2000) and rainbow trout exposed to 2.3 and 18 mg NP/l for 4 days (Shelley et al., 2012). Senthil Kumaran et al. (2011) reported alterations of hematocrit values in Clarias gariepinus exposed for 7 days to NP concentrations higher than those investigated in our study (Z250 mg/l). NP has been shown to affect circulating concentrations of several hormones. Zaccaroni et al. (2009) reported that thyroid hormones may serve as potential early biomarkers of NP endocrine disruption. In the present study, exposure to NP did not alter T4 levels in copper redhorse. Lerner et al. (2007) also reported no alteration of T4 levels in juvenile Atlantic salmon exposed to 10 and 100 mg NP/l for 21 days. However, other studies have shown a reduction of T4 levels in fish after intraperitoneal injection of NP (McCormick et al., 2005; Zaccaroni et al., 2009) or aqueous exposure to 50–100 mg/l NP (Sayed et al., 2012). The fact that many of the samples we analyzed had T4 levels below the detection limit of the assay may have prevented us from observing an effect of NP. Concentrations of T3 in copper redhorse exposed to up to 50 mg NP/l were similar to controls. Our results are in agreement with Lerner et al. (2007), Zaccaroni et al. (2009) as well as Sayed et al. (2012). The later observed a reduction in T3 levels in fish exposed to 80 and 100 mg NP/l but not at 50 mg NP/l. T and 11-KT are the major androgens in teleosts. Numerous studies have reported that NP and EE2 can affect sex steroids levels Table 2 Plasma concentrations of testosterone (T) and 11-ketotestosterone (11-KT), and brain aromatase activity, in immature copper redhorse exposed to NP and EE2 for 21 days. Data are expressed as means 7 S.E.M. (n ¼5–6). Treatment

T (pg/ml)

11-KT (pg/ ml)

Water control Solvent control 1 mg NP/l 10 mg NP/l 50 mg NP/l 10 ng EE2/l

115.4 74.4

73.6 7 23.0

1 3727 137

110.8 72.8

58.4 7 24.2

1 5067 157

51.6 7 21.1 33.4 7 11.8 15.8 7 2.1 28.8 7 9.3

1 4377 70 1 4947 89 1 6847 42 2 6197 231n

95.5 76.4 93.0 74.4 88.2 72.5n 87.9 76.3

Aromatase activity (fmol/h/mg protein)

n Significantly different from water and solvent controls (Kruskal–Wallis, po 0.01).


8000

**

7000

Plasma VTG (µg/ml)

6000 5000 4000 3000 2000 1000 0 Water

Solvent

1 µg NP/l

10 µg NP/l 50 µg NP/l 10 ng EE2/l

Treatment

4.5 *

Mucus VTG (µg/mg protein)

4 3.5 3 2.5 2 1.5 1 0.5 0 Water

Solvent

1 µg NP/l

10 µg NP/l 50 µg NP/l 10 ng EE2/l

Treatment

Fig. 1. (A) Plasma vitellogenin (VTG) concentrations and (B) surface mucus VTG levels in immature copper redhorse, before (white bars) and after (dark bars) a 21 day exposure to NP and EE2. Data are expressed as means 7 S.E.M (n¼ 5–6). The asterisks indicates a significant difference from both controls (Kruskal–Wallis, n ¼ p o 0.05, nn ¼ p o0.01).

Mucus VTG (µg/mg protein)

10

1

0.1

0.01 10

100

1000

10000

100000

Plasma VTG (µg/ml)

Fig. 2. Correlation between plasma vitellogenin (VTG) concentrations and surface mucus VTG levels of samples collected at the end of exposure.

173

in the plasma of fish (Schwaiger et al., 2002; Kortner et al., 2009; Salierno and Kane, 2009; Flores-Valverde et al., 2010; Sayed et al., 2012). The present work shows depressed T concentrations in immature copper redhorse exposed 50 mg NP/l. Average 11-KT concentrations were reduced after exposure to 50 mg NP/l, but the reduction was not significant. Depressed concentrations of T or 11KT were observed in other fish species exposed to NP (Yang et al., 2008; Sayed et al., 2012) or to effluents from treatment plants (Folmar et al., 1996; Walker et al., 2009). Depression in androgen levels mediated by NP could be due to alteration in steroidogenesis. Laurenzana et al. (2002) showed that, in the rat, NP inhibited the activity of enzymes involved in testosterone synthesis. Like NP, EE2 was shown to reduce T and 11-KT in fish plasma (Salierno and Kane, 2009; Flores-Valverde et al., 2010). In the present study, exposure to 10 ng EE2/l resulted in lower average levels of T and 11-KT, but these values were not significantly different from controls. Altered estradiol (E2) levels in plasma were observed in fish exposed to NP (Yang et al., 2008; Sayed et al., 2012) or EE2 (Salierno and Kane, 2009; Flores-Valverde et al., 2010). However, in the present work, plasmatic E2 concentrations could not be determined since they were below the detection limit, probably due to the immature status of the copper redhorse. The brain aromatase activity, in our study, was not significantly changed after exposure to NP for 21 days. Hallgren and Olsen (2010) also reported no effect of NP (10 and 50 mg/l) on the brain aromatase activity of male or female guppy after 2 week exposure. In contrast, EE2, 10 ng/l, produced a significant increase in aromatase activity in the brain of copper redhorse. The induction of brain aromatase activity by EE2 has also been shown in zebrafish (Andersen et al., 2003a), Japanese medaka (Oryzias latipes, Contractor et al., 2004) and guppy (Hallgren and Olsen, 2010). Other studies have reported that CYP19b mARN levels are increased after exposure to EE2 (Lyssimachou et al., 2006; Kortner et al., 2009) and E2 (Gelinas et al., 1998; Lee et al., 2000). The presence of estrogen responsive elements in the promoter region of the teleost CYP19b gene suggests that estrogenic compounds, such as EE2, might directly modulate the expression of CYP19b gene (Tchoudakova et al., 2001; Kuhl et al., 2005; Cheshenko et al., 2008). Our results support the idea of using aromatase activity or gene expression as a biomarker for exposure to EDCs (Meucci and Arukwe, 2006; Kortner et al., 2009; Diotel et al., 2010). VTG is a widely used biomarker of exposure to xenoestrogens. The capacity of NP and EE2 to induce VTG has been reported for numerous teleost species (Coady et al., 2010; Christen et al., 2010). However, the concentrations inducing VTG differs among studies (Coady et al., 2010; Christen et al., 2010). We observed a massive and significant VTG increase in plasma of copper redhorse exposed to 50 mg NP/l, while fish in other treatment groups, including the 10 mg NP/l group, had VTG levels similar to controls. This suggests that the effective concentration of NP for VTG induction in immature copper redhorse, under our experimental conditions, is between 10 and 50 mg NP/l. Our results are in agreement with recent studies showing that exposure to NP, at concentrations ranging from 3 to 100 mg/l, induces VTG in different fish species (reviewed in Coady et al., 2010). In the present work, VTG was not induced in copper redhorse exposed to EE2. Similarly, VTG was not induced in Japanese medaka exposed to 10 ng EE2/l (Örn et al., 2006). In contrast, many studies have reported an induction of VTG in fish exposed to EE2 concentrations lower than 10 ng/l (Jobling et al., 1996; Thorpe et al., 2003; Zha et al., 2007). The immature status of the copper redhorse is probably not the cause of the absence of VTG induction in the EE2 exposed group since VTG induction has been observed frequently in immature fish exposed to EDCs (Andersen et al., 2003b; Meucci and Arukwe, 2005; Arukwe and Røe, 2008). In addition, Tyler et al. (1999) have reported that VTG induction in fathead minnow (Pimephales

174


Fig. 4. Hepatic heat shock protein 70 (HSP70) expression in immature copper redhorse exposed to NP and EE2 for 21 days. The intensity of the HSP70 protein signal was determined by densitometry and expressed as relative values (means7 S.E.M, n ¼5–6). The relative values were defined as the intensity of signal of each sample divided by the intensity of signal of a reference sample added in each run. The asterisk indicates a significant difference from both controls (Kruskal–Wallis, p o0.01). (Inset) Representative Western blots of HSP70 in liver cytosolic fraction from the six treatment groups.

Fig. 3. (A) Hepatic microsomal cytochrome P4501A (CYP1A) protein expression in immature copper redhorse exposed to NP and EE2 for 21 days. The intensity of the CYP1A protein signal was determined by densitometry and expressed as relative values (means7 S.E.M, n ¼5–6). The relative values were defined as the intensity of signal of each sample divided by the intensity of signal of a reference sample added in each run. (Inset) Representative Western blots of CYP1A in liver microsomal fractions from the six treatment groups. (B) Hepatic microsomal ethoxyresorufinO-deethylase (EROD) activity in immature copper redhorse exposed to NP and EE2 for 21 days. Data are expressed as means 7S.E.M (n¼ 5–6). No statistical difference was observed between treatments (ANOVA, p 40.05).

promelas) at early life stages is at least as sensitive to estrogen as in adult fish. Experimental conditions during the exposure and a difference in species susceptibility to EDCs are factors that can contribute to differences in responsiveness observed among studies. Lange et al. (2012) demonstrated, using estrogen receptor in vitro assays and in vivo exposures, that the EE2-induced vitellogenic response differs among fish species, with carp, a cypriniform like the copper redhorse, being among the least responsive. These authors reported no hepatic VTG mRNA induction in carp following exposure to 10 ng EE2/l (nominal concentration) for 7 days. While VTG is routinely measured in plasma of fish, it can also be measured in the surface mucus (Gordon et al., 1984; Kishida and Specker, 1994). VTG has been detected in surface mucus of several species injected with, or exposed to, estrogenic chemicals (Kishida et al., 1992; Moncaut et al., 2003; Meucci and Arukwe, 2005; Rey Vázquez et al., 2009; Genovese et al., 2012) including copper redhorse and shorthead redhorse (Maltais and Roy, 2007,

2009). In a previous paper (Maltais et al., 2010), we demonstrated that levels of VTG in mucus from copper redhorse injected with ß-estradiol 3-benzoate were highly correlated with plasma levels. The physiological significance of the presence of VTG in the surface mucus of fish is unkown. It has been suggested that VTG in mucus is derived from the circulation and that skin could serve as an excretory pathway for excess plasma proteins (Kishida and Specker, 1994; Moncaut et al., 2003; Meucci and Arukwe, 2005). Moreover, presence of VTG mARN in epidermal cells of 4-NP treated Atlantic salmon suggests that the fish skin could be a site for synthesis of VTG (Arukwe and Røe, 2008). In the current study, we also observed a strong correlation between VTG levels in plasma and surface mucus. As in plasma, the exposure to 50 mg NP/l induced a significant increase of mucus VTG, while other treatments did not affect mucus VTG levels. Our results confirm the use of surface mucus sampling as a possible less-invasive alternative to plasma sampling for assessment of VTG in copper redhorse. The presence of EDCs in the habitat of the copper redhorse has been suspected to be a potential cause of the reproduction problems encountered by the species (Gendron and Branchaud, 1997). Sampling plasma may be too invasive or not allowed for an endangered species such as the copper redhorse, yet surface mucus still can be collected. Measurement of VTG in surface mucus of males or immatures from the field could be useful to investigate copper redhorse exposure to EDCs. In the present study, neither NP nor EE2 affected basal CYP1A expression or EROD activity in liver. A strong correlation between CYP1A protein levels and EROD activity was observed, suggesting no inhibition of EROD activity and no enzyme degradation or denaturation during sample preparation. Several studies have shown that hepatic CYP1A protein levels and its EROD activity in fish are reduced by NP (Arukwe et al., 1997; Vaccaro et al., 2005; Sturve et al., 2006; Cionna et al., 2006; Carrera et al., 2007) or EE2 (Solé et al., 2000b), presumably due to a crosstalk between the ER and AhR. Kirby et al. (2007) observed a suppression of dibenz[a,h] anthracene-induced hepatic EROD activity in flounder following waterborne exposure to NP and EE2. However, exposure to NP (100 mg/l) and EE2 (20 ng/l) did not affect basal hepatic EROD


activity in this species (Kirby et al., 2007), as observed in copper redhorse. Experimental design, as well as factors such as species, age, sex and reproduction stage, may influence the effects of estrogenic compounds on CYP1A levels and activity. HSP70, have been shown to be expressed in fish in response to toxicants, including EDCs. (Vijayan et al., 1998; Iwama et al., 1998; Aït-Aïssa et al., 2000; Carnevali and Maradonna, 2003). NP exposure resulted in an increase of hepatic HSP70 protein levels in copper redhorse (significant at 50 mg/l), which is in agreement with other studies. An increase in hepatic HSP70 mRNA levels was reported in Gobius niger injected with NP (Carnevali and Maradonna, 2003; Maradonna and Carnevali, 2007). Palermo et al. (2012) have reported an increase in HSP70 mRNA levels in the brain of juvenile sole, Solea solea, exposed for 7 days to 220 mg NP/l (10 6 M). Increased hepatic HSP70 mRNA levels were also observed in sole exposed to moderately contaminated sediment (Ribecco et al., 2012). Our results, combined with studies cited previously, support the idea of using HSP70 as a biomarker of environmental stress including exposure to exogenous pollutants. EE2 had no effect on hepatic HPS70 protein expression in copper redhorse. This is in accordance with Chandra et al. (2012) who reported that liver HSP70 protein expression was largely unaffected in mummichog (Fundulus heteroclitus) exposed to 50 and 250 ng EE2/l for 14 days. Information regarding the effects of NP on fish ChE activity is very limited. Our data indicate that exposure to NP did not affect AChE activity in the muscle of copper redhorse. Li (2008) reported a significant ChE inhibition in muscle of guppies exposed for 7 days to 60 and 150 mg/l of NP, concentrations greater than those tested in our study. AChE inhibition generally causes an increase of acetylcholine that may result in a decrease in the number of cholinergic receptors as a compensatory response to acetylcholine accumulation. Jones et al. (1998) reported a decrease in muscarinic cholinergic receptors in the brain of three species of trout exposed to NP for 96 h. However, this decrease was observed at NP concentrations equal to or above 50 mg/l and was not observed in fathead minnow and Colorado squawfish (Jones et al., 1998). More studies are needed to clarify the effects of NP on the ChE activity of fish. Greco et al. (2007) reported an increase of muscle AChE activity in Atlantic salmon exposed to EE2. However, this response was observed after 3 days exposure but had disappeared by day 7 suggesting that the fish had adapted to the EE2 exposure (Greco et al., 2007). This absence of response after a prolonged exposure is in accordance with our results since copper redhorse were exposed to EE2 for 21 days. In the present study, copper redhorse were exposed to waterborne NP and EE2 for 21 days under controlled conditions. NP induced VTG synthesis and hepatic HSP70 protein expression and altered testosterone levels, but had no effect on brain aromatase activity. In contrast, EE2 exposure increased brain aromatase activity but had no other significant effect. Both compounds have affinity for the ER but their effects on the biomarkers might be affected by other physiological aspects such as tissue distribution, metabolism and reproduction status. This support the use of multiple endpoints when asssessing the effects of EDCs on fish. Altogether, our results suggest that the copper redhorse is not particularly sensitive to a short-term exposure to NP and EE2 compared to other fish species. Significant effects of NP were observed at a high environmentally relevant concentration, i.e. 50 mg/l, and EE2, also at a high environmentally relevant concentration, had effect only on brain aromatase activity. However, in the wild, environmental conditions, diet, presence of other contaminants, long-term exposure and exposure at critical moments of development or reproduction may influence the toxicity of NP and EE2.

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5. Conclusion In conclusion, exposure of immature copper redhorse to environmentally relevant concentrations of NP and EE2 affected constituents of their endocrine system. This study suggests that NP and EE2 may have an impact on reproductive physiology in copper redhorse. Alteration of the endocrine system may also have deleterious impacts on growth, sexual differentiation and behaviors. Further research is needed to assess this. A priority of the provincial recovery plans and the federal recovery strategy for the copper redhorse is to evaluate exposure to EDCs and diagnose the possible role of EDCs in the difficulties facing copper redhorse reproduction (Équipe de rétablissement du chevalier cuivré, 2004, 2012; DFO, 2012). Measurement of non-invasive biomarkers, such as VTG in surface mucus, in wild copper redhorse, might bring additional information on their level of exposure to EDCs and reproductive difficulties, and contribute to restoration efforts of this endangered species by influencing research priorities to improve water quality and ensure conditions suitable for normal reproduction and growth.

Acknowledgments We are grateful to Yves Gauthier (Montreal Biodome) for the gift of copper redhorse. We thank Dr. Catherine Couillard for EROD analyses and for critical reading of the manuscript. We also thank Shirley Ye Guan, Véronique Desborbes, Benoit Legaré, France Boily and Mélanie Laflèche for technical assistance and Bernard Chenard for advice on fish culture. This work was financed by a matching grant from Fisheries and Oceans Canada, under the Species at Risk program. We thank the two anonymous reviewers whose comments helped to improve the manuscript. Domynick Maltais would like to acknowledge the mentorship and friendship of Dr. Robert Roy who left us at the beginning of 2014.

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Comparative physiology and relative swimming performance of three redhorse (Moxostoma spp.) species: associations with fishway passage success.

Reviewer: an endangered species?!

Pubic lice: an endangered species?

The clinical investigator as an endangered species.

Ideas are becoming an endangered species.

Cumulative profile: rural FPs: an endangered species?

Endangered species.

Inbreeding effects in Solanum lycocarpum A. St.-Hil populations, an endangered species of the Brazilian Cerrado.

Micropropagation of Sterculia urens Roxb. - an endangered tree species.

Effects of Lead on Ultrastructure of Isoetes sinensis Palmer (Isoetaceae), a Critically Endangered Species in China.

The physician scientist: an endangered but far from extinct species.

A Mixed-Method Approach for Quantifying Illegal Fishing and Its Impact on an Endangered Fish Species.

Endangered species. Emerging virus threat.

The Effects of p-nonylphenol on the Myometrial Contractile Activity.

Sublethal toxic effects of nonylphenol ethoxylate and nonylphenol to Moina macrocopa.

Estimating effects of tidal power projects and climate change on threatened and endangered marine species and their food web.

Endangered Species. Red wolves in the crosshairs.

Bioeconomic analysis supports the endangered species act.

Genetic structure of colline and montane populations of an endangered plant species.

India: Endangered species damned by dams.

Fluctuating fire regimes and their historical effects on genetic variation in an endangered shrubland specialist.

Books and other endangered species: an inquiry into the values of medical librarianship.

Ecological risk assessment of nonylphenol in coastal waters of China based on species sensitivity distribution model.

Effects of copper sulfate, copper sulfide and sodium sulfide on performance and copper stores of pigs.