Use of Early Juvenile Zebrafish Danio Rerio for In-Vivo Assessment of Endocrine Modulation by Xenoestrogens

This is an electronic reprint of the original article. This reprint may differ from the original in pagination and typographic detail. Author(s): Title: Year: Version: Please cite the original version: All material supplied via JYX is protected by copyright and other intellectual property rights, and duplication or sale of all or part of any of the repository collections is not permitted, except that material may be duplicated by you for your research use or educational purposes in electronic or print form. You must obtain permission for any other use. Electronic or print copies may not be offered, whether for sale or otherwise to anyone who is not an authorised user. Use of Early Juvenile Zebrafish Danio Rerio for In-Vivo Assessment of Endocrine Modulation by Xenoestrogens Sahoo, Tarini Prasad; Oikari, Aimo Sahoo, T. P., & Oikari, A. (2013). Use of Early Juvenile Zebrafish Danio Rerio for In- Vivo Assessment of Endocrine Modulation by Xenoestrogens. Journal of Environmental and Analytical Toxicology, 4(1), 1-13. https://doi.org/10.4172/2161- 0525.1000202 2013 ISSN: 2161-0525 Journal of Environmental & Analytical Toxicology The International Open Access Journal of Environmental & Analytical Toxicology Executive Editors Manuel Lerdau University of Virginia, USA Warren G. Foster McMaster University, Canada Lygia T. Budnik University Medical Centre Hamburg, Germany Ben-Zhan (Benny) Zhu State Key Laboratory of Environmental Chemistry and Ecotoxicology, China Aijie Wang Harbin Institute of Technology, China This article was originally published in a journal by OMICS Publishing Group, and the attached copy is provided by OMICS Publishing Group for the author’s benefit and for the benefit of the author’s institution, for commercial/research/educational use including without limitation use in instruction at your institution, sending it to specific colleagues that you know, and providing a copy to your institution’s administrator. All other uses, reproduction and distribution, including without limitation commercial reprints, selling or licensing copies or access, or posting on open internet sites, your personal or institution’s website or repository, are requested to cite properly. Available online at: OMICS Publishing Group (www.omicsonline.org) Digital Object Identifier: http://dx.doi.org/10.4172/2161-0525.1000202 Research Article Open Access Volume 4 • Issue 1 • 1000202 J Environ Anal Toxicol ISSN: 2161-0525 JEAT, an open access journal Open Access Research Article Environmental & Analytical Toxicology Sahoo and Oikari, J Environ Anal Toxicol 2013, 4:1 http://dx.doi.org/10.4172/2161-0525.1000202 *Corresponding author: Tarini Prasad Sahoo, Division of Environmental Science and Technology, University of Jyväskylä, Jyväskylä, 40014, Finland, Tel: 358408053900; Fax : 358142602321 ; E-mail: [email protected] Received October 21, 2013; Accepted December 24, 2013; Published December 27, 2013 Citation: Sahoo TP, Oikari A (2013) Use of Early Juvenile Zebrafish Danio Rerio for In-Vivo Assessment of Endocrine Modulation by Xenoestrogens. J Environ Anal Toxicol 4: 202. doi: 10.4172/2161-0525.1000202 Copyright: © 2013 Sahoo TP, et al. This is an open-access article distributed under the terms of the Creative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original author and source are credited. Use of Early Juvenile Zebrafish Danio rerio for In-Vivo Assessment of Endocrine Modulation by Xenoestrogens Tarini P Sahoo* and A Oikari Division of Environmental Science and Technology, University of Jyväskylä, Jyväskylä, Finland Keywords: Brain aromatase; Early-life stage; Vitellogenin1; Xenoestrogenicity; Zebrafish Introduction Diverse groups of natural and synthetic compounds known to have an endocrine bioactive potential are constantly introduced into the environment. These endocrine disrupting compounds (EDCs) have been found to exhibit significant effects on reproduction by acting on endocrine functions [1,2]. Given the extent of organic pollutants released into aquatic environment, teleosts among vertebrate taxa qualify as candidates to investigate endocrine modulation [2-6]. The choice of zebrafish Danio rerio (Hamilton, 1822) as a toxicological model presents practical advantages, including cost-effective maintenance, asynchronous breeding, and rapid development of the embryo-larval stages. Moreover, the sexual differentiation in the brain determining the fate of the gonads in teleosts, contrasts with amphibians and mammals [7]. In general, signaling cues in the form of endogenous hormones and external factors stimulate the pituitary to release gonadotropins, which in turn stimulate the gonads to synthesize and release endogenous hormones, sex steroids. Under endocrine control, the hypothalamuspitutary-gonadal (HPG) axis is further extended to the liver (HPGL), the site for synthesis of the precursor of yolk protein, vitellogenin (Vtg), common to oviparous vertebrates. The hepatic synthesis of these proteins is stimulated by the binding of physiological estrogens or xenoestrogens to estrogen receptors (ERs) in the liver (Figure 1). In zebrafish, seven isoforms of vtg have been identified, with dominant expression of vtg1 compared to vtg2-7 [8]. Males possessing the normally unexpressed vtg genes when exposed to exogenous estrogen or their mimics, up-regulate vitellogenesis [1]. Consequently, alteration of endocrine signaling along the HPGL axis and subsequent modulation of gametogenesis and other reproductive processes from teleost responses indicates mechanisms of EDC action [9]. Bio-chemical pathways have been used to link health and reproductive status in wildlife to environmental chemical exposures. Steroid biosynthetic enzymes catalyze the conversion of cholesterol into active sex hormones (estradiol and testosterone) mainly in brain, kidney and gonads [10]. Key enzymes in steroidogenesis include steroidogenic acute regulatory (Star) protein, cytochrome P450 side-chain cleavage (P450scc) or Cyp11a1, 3-beta-hydroxysteroid dehydrogenase (Hsd3b1), and cytochrome P450 aromatase (P450arom) or Cyp19a1 (Figure 1). Other studies have also discussed regulation in the steroidogenic pathway with gene expression analyses with adult fathead minnow [11] and zebrafish (in vitro, [12]; in vivo, [13,14]. Immunohistochemical studies showed the significance of (endogenous) sex steroids on gonadal sex differentiation in teleosts [9,15]. Other studies did not support the involvement of endogenous steroids in the gonadal sex differentiation in fish species such as medaka [16]. Therefore, it was interesting to compare in this study the expression patterns of key steroidogenic enzymes in early juvenile zebrafish, if any. Importantly, the synthesis and release of sex steroids is controlled by the aromatization of C19 androgens (testosterone) to C18 estrogens (17β-estradiol) by aromatase. In teleosts, the expression of tissuespecific isoforms of P450arom genes (gonad, cyp19a1a; brain, cyp19a1b) [17], with significantly higher relative expression of the brain isoform, has been used to indicate estrogenic stimulation of sex steroid synthesis [18-23]. Besides induction of vtg1, alterations of steroid biosynthesis and plasma steroids are biomarkers indicative of estrogenic action [24]. Abstract Reliable and cost-effective early-life stage (ELS) bioassays incorporating practical experimentation without compromising scientific relevance are crucial in chemical risk assessment. This study investigated the use of 20 dayspost-fertilization life stage (20dpfZF) of zebrafish Danio rerio to screen environmental chemicals known to be estrogenic in adult fish. Firstly, studies with key genes in steroidogenesis were conducted; the brain isoform of aromatase gene (cyp19a1b) being the most prominently expressed biomarker. Regulation of mRNA levels of molecular biomarkers, vitellogenin 1 gene (vtg1) and cyp19a1b were selected to assess the endocrine modulation by xenoestrogens, 17α-ethinylestradiol (EE2), 4-n-nonylphenol (NP), 4-t-octylphenol (OP) and bisphenol A (BPA). Groups of 20dpfZF (n=15) as three replicates were exposed to chemicals over a five-day period in aerated static setups. Exposure of 20dpfZF to sediment spiked with EE2 (nominal 3µg g-1dw) was also conducted to assess the sensitivity of this life-stage to sediment with estrogenic potency. Whole body homogenates of exposed juveniles showed the estrogenic potential of chemicals in the order: EE2 > OP > BPA > NP. Higher relative expression of cyp19a1b was noticed at lower ambient concentrations of EE2, although vtg1 showed more pronounced expression to it. The 20dpfZF responded in a dose-related way to sediment spiked with EE2, expanding its use as a general aquatic animal model. The suitability of 20dpfZF as an in vivo model, along with stable expression of reference genes was established. In addition to consistent expression pattern of key target genes on xenoestrogenicity, it serves as a practical screening model for the risk assessment of environmental chemicals and samples with estrogenic potential. Citation: Sahoo TP, Oikari A (2013) Use of Early Juvenile Zebrafish Danio Rerio for In-Vivo Assessment of Endocrine Modulation by Xenoestrogens. J Environ Anal Toxicol 4: 202. doi: 10.4172/2161-0525.1000202 Page 2 of 13 Volume 4 • Issue 1 • 1000202 J Environ Anal Toxicol ISSN: 2161-0525 JEAT, an open access journal Along with vtg1, gene expression of cyp19a1b, a central steroidogenic enzyme, present suitable molecular markers to detect and analyze the impact of estrogenic exposure, considering the short ovarian cycles and estrogen-responsive embryonic stages in zebrafish. The postembryonic life-stage (20 days-post-fertilization zebrafish; 20dpfZF) was used for a five-day exposure to 17α-ethinylestradiol (EE2), a common benchmark of xenoestrogens. Although ranging in nano-and micromolar concentrations in ambient water (0.2-40.0 ng L-1; [25]) and sediment (< 0.5-22.8 µg kg-1 dry mass; [26]) sources respectively, EE2 is a potent xenoestrogen known to have adverse effects on the endocrine function of aquatic organisms. The objective of this study was to evaluate a life-stage incorporating practical and experimental advantages for screening environmental estrogens, assessed by gene expression analyses. Pilot exposures to EE2 were carried out to select the key steroidogenic enzyme, cyp19a1b to complement vtg1 as a panel of estrogen-responsive biomarkers for subsequent studies. Besides EE2, the sensitivity of this life-stage was evaluated with water-borne exposures of other xenoestrogens with varying potencies-bisphenol A (BPA), 4–n–nonylphenol (NP) and 4-tert-octylphenol (OP). The screening potential of 20dpfZF was also verified with exposure to sediment spiked with EE2, and is discussed in context to regulatory requirements with the use of adult animals in toxicological research. Materials and Methods Chemicals and exposure stock solution 2,2-bis(4-hydroxyphenyl)propane (BPA, purity 99+%), 17α-ethynyl,3,5(10)-estratriene-3, 17β-diol (EE2, minimum 98% by HPLC), and 4-(1,1,3,3-tetramethylbutyl)phenol (OP, 97%) were purchased from Sigma-Aldrich (St. Louis, MO, USA). 4-n-Nonylphenol (NP, 98+ %) was obtained from Alfa Aesar (Karlsruhe, Germany). Silylation reagent N, O-bis(trimethylsilyl) trifluoroacetamide (BSTFA) with 1% trimethylchlorosilane (TMCS) was purchased from Fluka Chemie (Buchs, Switzerland). Internal standards 17-α-ethynylestradiol-d4 (EE2-d4, 97-98%) and 2,2-bis(4- hydroxyphenyl) propane-d16 (BPA-d16, 98%) were purchased from LGC Standards AB (Borås, Sweden). The solvents, methanol (MeOH) and acetonitrile, supplied by Merck Chemicals (Darmstadt, Germany) were of HPLC grade. Separate chemicals were dissolved in methanol, and stocks were stored at -20°C until use, within two weeks. Experimental animals Adult male and female zebrafish (Danio rerio) (wild type; WT) were obtained from Institute of Medical Technology, University of Tampere (Tampere, Finland). Fish were acclimatized to temperatures (mean ± standard deviation, 26 ± 1°C) and photoperiod (14 h light/10 h dark) for at least two weeks prior to setup for breeding. Animal husbandry was carried out along established guidelines (Westerfield, 2007). Zebrafish adults were fed twice daily with frozen chironomid larvae (Ruto frozen fish food, Zevenhuizen, Holland). Following breeding between sexually mature adults (~8 month old), fertilized eggs (0.5- 1.5 hours-post-fertilization; hpf) were collected and reared through early post-embryonic stages for 20 days (20dpfZF) in rearing water or embryo medium (Westerfield, 2007). Post-hatch zebrafish from 6 dpf through 25 dpf were fed with fry food (SDS diets, UK). Figure 1: An outline of the steroid biosynthetic pathway and the action of estrogen on vitellogenesis in fish and other vertebrates. The conversion of cholesterol to pregnenolone and subsequent steps occurs in the inner mitochondrial membrane. Homodimerization of ERs and binding to estrogen-responsive elements (ERE) in the regulatory region (promoter) of susceptible genes lead to transcription of estrogen-responsive genes. Citation: Sahoo TP, Oikari A (2013) Use of Early Juvenile Zebrafish Danio Rerio for In-Vivo Assessment of Endocrine Modulation by Xenoestrogens. J Environ Anal Toxicol 4: 202. doi: 10.4172/2161-0525.1000202 Page 3 of 13 Volume 4 • Issue 1 • 1000202 J Environ Anal Toxicol ISSN: 2161-0525 JEAT, an open access journal Chemical exposure and sampling Chemical exposures were conducted in glass jars containing 250 mL of exposure medium, consisted of gently aerated embryo-rearing water (pH 7.2) for the duration of the experiment, to avoid change in water quality. Water-borne exposures (I and II) were carried out in a semi-static manner, replacing 100% of the medium daily. Water quality parameters were measured on days 1 and 4. The mean temperature in the exposure jars was in the range 24-26°C. The content of dissolved oxygen (DO, mg L-1) was 7.6 ± 0.2 (mean DO ± sd) and pH 7.2 ± 0.2 (mean ± sd). Chemicals dissolved in the carrier solvent, methanol (MeOH), were administered to the exposure water, the solvent concentration being 0.02% or less. Preliminary exposures also included water controls (data not shown), i.e., without solvent to confirm similarity with that of the maximum solvent concentration (0.02%). The first set of exposures (Exposure I) were performed with groups of 30 fish (20dpfZF), replicated three times in parallel, exposed to 0.02% MeOH with 0, 5, 25 and 50 ng EE2 L-1, to determine expression and selection of suitable of steroidogenic gene(s). The second set of exposures (Exposure II) were performed with groups of 25 fish (20dpfZF), replicated three times, exposed to 0.02% MeOH with 0, 5, 25, and 50 ng EE2 L-1, 0, 100, 500, and 1000 µg BPA L-1, 0, 10, 50, and 100 µg NP L-1, and 0, 10, 50, and 100 µg OP L-1. Although the results used in discussion represent measurement from three parallel replicates, the pattern of relative expression of target genes were also verified with two independent exposures conducted earlier. Fish were sampled at the end of five days of exposure. Each group of sampled animals (n = 15) was collectively weighed in rnasefree micro centrifuge tubes (Starlab, Germany), and snap-frozen in liquid nitrogen and stored at -80°C until further analysis. Experiments were conducted according to and licensed by the Finnish authority for animal experiments (ESAVI-2010-07885/Ym-23). Sediment exposure The reference sediment (0-10 cm depth; dry weight, dw 5.2%; total organic content, TOC 4.7%) was sampled from Lake Palosjärvi (located in Central Finland) using an Eckman grab sampler. Chemical spiking of it was performed according to standard guidelines for wholesediment toxicity testing [27,28]. Briefly, nominal concentration 3 µg g-1 dw of EE2-spiked sediment was prepared by gradually administering EE2 (dissolved in 100% methanol) to sediment slurry with minimum amount of water mixed for ~7 h. Following the addition of EE2, the headspace of sample container was filled with an inert gas (N2) and sealed for equilibration at 4°C for a minimum of 30 d before the assays. The exposure series (Exposure III) were prepared in 600 mL glass beakers in a ratio 1:4 (dw/v) of sediment and overlying water (3°dH or 54 mg CaCO3 L-1), added slowly to avoid any suspension of particulates to the water column. The exposure beakers were allowed to stand for ~16 h at 24-26°C before the experimental fish (n=25) were added. The exposure series consisted of control sediment (Ref-Sed; i.e., the reference Palosjärvi sample) compared with EE2-spiked samples; Sed 1 (3 d), and Sed 2 (5 d). The exposures of animals were conducted in aerated static systems (no water renewal), sampled at the end of 3 d for Ref-Sed, Water-EE2 (150 ng L-1), Sed 1+ Salt and Sed 1, and at the end of 5 d for Sed 2 (Figure 4). Two controls treatments, each accompanying the respective 3- and 5 d exposures were included to provide for relative quantification per sampling time point. To maintain the water quality, the fish were not fed during the experiment. To assess the possible role of the influence of electrolyte content of overlying water, another sediment exposure Sed1 + Salt included E3 salts (embryo medium; 6°dH or 107 mg CaCO3 L-1) at levels used for rearing of embryo-larval stages of zebrafishup to 20dpfZF. For comparison of estrogenic effects, waterborne exposure of 20dpfZF to 150 ng EE2 L-1 for 3d was also included as positive control. Experimental conditions and fish sampling (n=3 pools of at least 15 fish each) was the same as in water-borne exposures. Solid-phase extraction (SPE) of water samples Ethinylestradiol was extracted from water samples by a SPE method [29] modified for the analyses in the present study (see supplementary information). Similar extraction was employed with NP and OP samples (3-10 mL) using BPA-d16 as the internal standard. GC-MS analysis of EE2 The gas chromatography–mass spectrometry (GC-MS) was performed with HP 6890 gas chromatograph (Hewlett-Packard, Walbronn, Germany) equipped with HP 5973 mass selective detector (Hewlett-Packard, Palo Alto, CA, USA). The method included a derivatization [30] by silylation with BSTFA, and pyridine to increase the stability of derivatized products and improve extraction recoveries of the target compounds (see supplementary information). Concentration of EE2 was calculated using relative spectral peak-areas of analytes, internal standard, and the determined response factor. Specifically, the results were based on determination of di-TMS derivatives of EE2 and EE2-d4. LC-MS /MS analysis of BPA, NP, and OP The analysis was performed with Waters Alliance 2795 (MA, USA) LC and the determination of target compounds was performed in negative ion electro spray mode (ESI-) and a Quattro Micro triplequadrupole mass spectrometer (MS/MS) (Waters, MA, USA) with electro spray interface was used as detector. The method described by Revilla-Ruiz et al. [31], with minor modifications, was conducted with extracted (NP and OP) and un-extracted (BPA) water samples. Data acquisition was performed with multiple reactions monitoring (MRM) mode and the corresponding parameters for the target compounds are presented in supplementary (Table SI). Isolation of RNA and reverse transcription Total RNA was extracted from homogenized pooled fish samples (15 animals with total weight ca. 16 mg) using the Tri Reagent (Molecular Research Center, Inc.) according to manufacturer’s instructions, and the quality and quantity were determined on a NanoDrop ND-1000 UV-visible spectrophotometer. Absorbance measurements including 260:280 ratios (1.8 to 2.0) and the additional 260:230 > 260:280 ratios indicated a secondary measure of nucleic acid purity [32]. The integrity of RNA was also verified using the Agilent RNA 6000 Nano Assay Protocol with an Agilent 2100 Bioanalyzer (Waldbronn, Germany). RNA integrity number (RIN, 1-10) generated for analyzed samples was checked within range 7-10 for all samples. The total RNA extracted was reverse transcribed into cDNA using iScript cDNA synthesis kit (Bio- Rad, USA), and the cDNA templates were diluted ten times for use in real-time PCR assays. Real-time PCR Real-time PCR was performed using iQ SYBR Green supermix (Bio-Rad, USA) and run on a Bio-Rad CFX96™ real-time PCR detection system (Bio-Rad, USA). Gene-specific primers (Table SII) for housekeeping genes (beta-actin (β-actin) and elongation factor 1 alpha (ef1a)) as well as target genes (star, cyp11a1, hsd3b1, cyp19a1b) were designed using Amplifx (v 1.5.4), and the given oligonucleotides (desalt purified) were purchased from Sigma-Aldrich (UK). The primer pair sequences were checked with the program Mfold (http://mfold.bioinfo.rpi.edu/ Citation: Sahoo TP, Oikari A (2013) Use of Early Juvenile Zebrafish Danio Rerio for In-Vivo Assessment of Endocrine Modulation by Xenoestrogens. J Environ Anal Toxicol 4: 202. doi: 10.4172/2161-0525.1000202 Page 4 of 13 Volume 4 • Issue 1 • 1000202 J Environ Anal Toxicol ISSN: 2161-0525 JEAT, an open access journal cgi-bin/dna-form1.cgi) to predict the potential for amplicon secondary structures that may prevent efficient amplification. To evaluate qPCR efficiency, amplification efficiency with standard curve analysis (PCR efficiency, 90-110%; slope, -3.1 to -3.6; R2 > 0.98) was performed, and optimization of qPCR reactions and protocol was carried out. The qPCR reaction mix was optimized for a 25 µL reaction mix consisting of 12.5 µL of SYBR green supermix, 2.5 µL of each of the primer pairs (200 nM), 2.5 µL sterile water, and 5 µL of cDNA template (100 ng µL–1). The PCR protocol comprised 10 min polymerase activation at 95°C, followed by 40 cycles of 30 sec at 95°C, and 30 sec at 60°C. Each sample was assayed in duplicate, and assays requiring multiple runs included inter-run calibrators (IRCs) on each of the plates to facilitate sample maximization and restricting reference gene reactions to a single plate. The inter-run calibration was performed to correct for variation, allowing comparison between plates within the same analysis. Data analysis The expressions of target genes were normalized to the expression of two reference genes, (β-actin, ef1a). The choice of reference genes was based on their stable expression across samples of different treatment. The determined stability parameter (M value) for the reference genes were used to select suitable reference genes with M values < 0.5; the lower the M-value, that higher the stability [33]. The relative gene expression levels were calculated from the experimental amplification data (Cq values; quantification cycles) using the 2-ΔΔCT method [34]. Statistical significance (P< 0.05) of log-transformed data from Exposures I and III was determined by a one-way analysis of variance (ANOVA) followed by Tukey HSD post hoc test. Data from Exposure II was analyzed using Welch’s test followed by Games-Howell post hoc test. All analyses were performed using IBM SPSS (v 19). Results Water quality and concentrations of xenoestrogens in experiments No significant difference in water quality was found between exposure jars or treatments during the exposure period. Regarding to nominal concentration added in experimental waters, actual chemical concentrations (mean ± sd) are presented in Table SIII. However, all measured concentrations of NP were not detected. Since both NP and OP samples were extracted simultaneously with the same method, the undetected NP samples may have resulted from issues with the extraction method. For clarity, while the measured concentrations commonly ranged 20-50% below expected, revealing effects, the nominal values are used in further discussion. Exposure I: Responses of steroidogenic enzyme mRNA levels The mean M-values of reference genes, β-actin and ef1a across the corresponding concentration series of exposures (BPA, NP, OP, and EE2) were 0.004, 0.1, 0.3, 0.06; and 0.01, 0.04, 0.06, 0.09 respectively, showing stable expression of the respective housekeeping genes. For normalization of gene expression, the mean M-values of both reference genes combined and considered together for the chemical exposures were 0.01, 0.07, 0.06, and 0.07 respectively, suggesting the stable expression of this reference gene pair across chemical treatment at the studied life-stage. Figure 2 shows the normalized gene expression levels of cyp11a1, cyp19a1b, hsd3b1, and star in 20dpfZF exposed to EE2 for five days (20- 25 dpf). Among the studied genes, only cyp19a1b showed pronounced expression levels compared with cyp11a1, hsd3b1, and star, (Tukey HSD, P< 0.001). Noticeable induction of cyp19a1b resulted formal EE2 exposures (5, 25 and 50 ng L-1), averaging 13.5, 18.7 and 22.8-fold changes respectively, with gradual increase with increasing concentrations (5-50 ng L-1). In preliminary studies with similar experimental conditions, cyp19a1b showed ~2-fold increase at even lower exposures (1 ng L-1) (data not shown). There was no induction of star in exposure groups, with expression even in the highest EE2 exposure, 50 ng EE2 L-1, being similar with that of control and 5 ng EE2 L-1. Although 25 ng EE2 L-1 revealed ~1.3-fold change, the altered star expression was not significantly different from either the control or respective expression levels of cyp11a1 and hsd3b1 (Tukey HSD, P > 0.05). Similarly, no statistically significant effects were observed with the mRNA levels of cyp11a1 and hsd3b1 that actually showed decreased average expression compared to solvent control. Overall, there was a statistically significant difference between expression of cyp19a1b and the other steroidogenic enzymes (star, cyp11a1 and hsd3b1) as determined by one-way ANOVA for EE2 exposures 5 ng L-1 (F(3, 8) = 91.740, P< 0.001), 25 ng L-1 (F(3, 8) = 61.130, P< 0.001) and 50 ng L-1 (F(3, 8) = 133.082, P< 0.001). A Tukey HSD post hoc test revealed the significant up-regulation of only cyp19a1b expression among the studied steroidogenic enzymes (P< 0.001) compared to star–cyp11a1 (P = 0.848) and star–hsd3b1 (P = 0.982). Exposure II: Changes in cyp19a1b and vtg1 mRNA levels to xenoestrogens in water Figure 3a shows the relative expression of cyp19a1b and vtg1 in 20dpfZF exposed to EE2 from 20 through 25 dpf. The mRNA levels of cyp19a1b exposed to 5, 25, and 50 ng EE2 L-1 were induced to 13, 19, and 24-fold respectively, relative to control treatment. For the same concentrations, relative expression levels of vtg1 were 5, 143, and 1272- fold. Despite the larger extent of vtg1 expression across the exposure treatment, cyp19a1b showed more sensitive induction (13-fold; Games- Howell, P< 0.05) at the lowest range (5ng EE2 L-1) compared to vtg1 (5- fold; Games-Howell, P< 0.05). As presumed, the pattern of expression of cyp19a1b at all concentrations of EE2 was similar to samples studied for steroidogenic enzyme mRNA levels (Figure 2d); with gradual increase from 5 to 50 ng EE2 L-1. However, at higher treatments levels (25 and 50 ng EE2 L-1), the relative induction of vtg1 (Games-Howell, P< 0.01, 0.001) was noticeably higher than cyp19a1b (Games-Howell, P< 0.05), albeit both the genes showing significant relative expression. Almost 15% mortality was observed in groups exposed to the highest nominal concentration of EE2 (Figure S1). Whole-body cyp19a1b showed increased expression levels at 100, 500 and 1000 µg BPA L-1 compared to vtg1 (Figure 3b). Expression of vtg1 was down-regulated (in average -1.2-fold change) by lower concentrations of BPA (100 and 500 µg L-1), but was significantly upregulated at the highest concentration (2.7-fold change; Games-Howell, P< 0.05). The two higher concentrations of BPA L-1 showed significant induction of cyp19a1b transcripts, 2.9 and 4.4-fold (Games-Howell, P< 0.05) respectively. Clearly, cyp19a1b was more sensitive compared to vtg1 across a range of BPA exposures, further revealing that BPA induces a weak estrogenic effect when compared to EE2. In 20dpfZF exposed to 10, 50 and 100 µg NP L-1, the relative expression levels of both cyp19a1b and vtg1 showed no statistically significant change (Games-Howell, P> 0.05), showing a clear downregulation for both genes (Figure 3b). Compared to NP, the expression pattern was different with OP exposure; cyp19a1b showing an upregulatory pattern for all three concentrations compared to control (Figure 3b). The relative induction of vtg1 was significant only at 100 Citation: Sahoo TP, Oikari A (2013) Use of Early Juvenile Zebrafish Danio Rerio for In-Vivo Assessment of Endocrine Modulation by Xenoestrogens. J Environ Anal Toxicol 4: 202. doi: 10.4172/2161-0525.1000202 Page 5 of 13 Volume 4 • Issue 1 • 1000202 J Environ Anal Toxicol ISSN: 2161-0525 JEAT, an open access journal µg OP L-1 (2.5-fold; Games-Howell, P< 0.05). For all the concentrations, cyp19a1b again showed higher sensitivity for the induction, i.e., 1.9, 1.7, and 2.5-fold. With OP, as in the case of EE2 and BPA, cyp19a1b was a more sensitive biomarker of estrogenicity than vtg1. Overall, it is noteworthy that NP compared to EE2, BPA, and OP had a much less estrogenic effect in the 20dpfZF model. There was statistically significant difference between exposure groups determined by Welch’s test for both cyp19a1b (F (11, 9.197) = 422.008, P< 0.001) and vtg1 (F (11, 9.355) = 666.860, P< 0.001). Both 50 and 100 µg OP L-1 affected the survival of exposed fish to ~20 and 30% mortality (Figure S1). Exposure III: Changes in cyp19a1b and vtg1 mRNA levels to sediment spiked with EE2 In relevant contexts of in vivo exposures, 20dpfZF showed preferential use of water-column depending on exposure type (with and without sediment). While each group of 25 animals in the wateronly units appeared to distribute randomly across the water column, the fish in units with sediment headed for the sediment-water interface, perhaps seeking food. Actually, by actively ingesting contaminated sediment, the possibility to study diverse exposure scenarios with 20dpfZF (i.e. contaminants taken up by ingestion) was obvious (Figure 5b). We suggest that the same is not possible with sessile and nonfeeding life stages. One-way ANOVA analysis of normalized gene expression showed significant induction of both cyp19a1b (F (4, 10) = 1096.501, P< 0.001) and vtg1 (F (4, 10) = 2471.846, P< 0.001) of 20dpfZF when exposed to sediment spiked with nominal 3 µg EE2 g-1dw over three to five days (Figure 4a, 4b). Between the exposures (Sed1 and Sed2) conducted over three and five days, Tukey’s HSD post hoc test showed that fish exposed to sediments with EE2 exhibited higher induction of cyp19a1b than the water-borne exposure (nominal 150 ng EE2 L-1) at 63- and 50- fold change, respectively (Figure 4a). Significant up-regulation of Sed 2 fish with 74-fold change, compared to EE2-water exposure, indicated the influence of the extended exposure duration. For the 3d exposures (Sed1+Salt, Sed1), which differed only in respect to the electrolyte content, showed higher induction (72-fold) compared to that of Sed1, although not significantly different (Tukey’s HSD, P = 0.473). The expression of vtg1 showed very pronounced up-regulation patterns, increasing from EE2 in water (181-fold) to Sed 1 (500-fold) and Sed 2 (1152-fold) (Figure 4b). Unlike cyp19a1b, vtg1 showed significantly lower induction in Sed 1 + Salt (198-fold) when compared to Sed 1 (500-fold), possibly as an indication of the difference in the fraction of EE2 taken up by the fish (Tukey’s HSD, P< 0.001). These may be considered consistent with differences in the partitioning and bioavailability of EE2 between static water and sediment exposures systems, based on ambient characteristics. Figure 2: Exposure I: Changes in mRNA levels of steroidogenic enzymes. Gene expression of star (a), cyp11a1 (b), hsd3b1 (c) and cyp19a1b (d) in 20dpfZF exposed to concentrations of EE2 for five days (from 20 to 25 dpf). Box plots represent median (—), 25th and 75th percentiles (lower and upper end of boxes), minimum and maximum values (end of the whiskers) for polled whole-body homogenates (n=3 pools of at least 20 fish each). Target genes were normalized to geometric mean of two reference genes (β-actin, ef1a). Nominal exposure concentrations include 0, 5, 25 and 50 ng EE2 L-1. Means with different letters are significantly different (P< 0.05) as determined by One-way ANOVA followed by Tukey’s HSD post hoc test. Citation: Sahoo TP, Oikari A (2013) Use of Early Juvenile Zebrafish Danio Rerio for In-Vivo Assessment of Endocrine Modulation by Xenoestrogens. J Environ Anal Toxicol 4: 202. doi: 10.4172/2161-0525.1000202 Page 6 of 13 Volume 4 • Issue 1 • 1000202 J Environ Anal Toxicol ISSN: 2161-0525 JEAT, an open access journal Discussion Among aquatic vertebrates, the zebrafish has been widely applied in investigating biological responses in development [35], endocrine disruption [3,4,36], and reproduction [37] among other effects. As adult organisms, the fish has been established as a reliable and sensitive organism in studies of comparative toxicology. On the other hand, in line with alternative animal testing requirements, ELS bioassays are aimed to substitute more advanced stages for toxicity testing [37-39]. The feasibility to expose ELS includes a large sample size, not possible with adult fish. Moreover, an early in vivo model presents an integrated organism system compliant with regulatory mandates of vertebrate animal testing. The reliability of gene level responses [40,41], including estrogenic and the other steroidogenic ones, is suggested to give a mechanistic understanding on impacts of environmental chemicals, their dynamics of effects, and most importantly, extrapolations used in environmental risk assessment. Introduction of the 20dpfZF clearly satisfied several scientific, practical, and cost-effective requirements, although the main focus of this study was to select and evaluate the life-stage as a suitable screening model for xenoestrogenicity. The information regarding the gene-level response determined here in water-based exposures could be compared to a sediment exposure carrying a potent xenoestrogen with substantial trait of hydrophobicity, EE2 (log Kow 4.15) [42]. Exposure of zebrafish (20-25 dpf) to EE2, BPA, NP, and OP was used to evaluate regulation of cyp19a1b and vtg1 transcript levels. The choice of cyp19a1b as a core endpoint for this life-stage was determined from concentration-specific transcript abundance compared to other steroidogenic enzymes (star, cyp11a1 and hsd3b1; Figure 1). The presence of ERE in the promoter region of cyp19a1b, but absent on cyp19a1a, has been shown to explain the estrogen-responsiveness of cyp19a1b [7,18,43]. Exposure I: Selection of steroidogenic endpoint for xenoestrogenicity The steroid biosynthetic pathway can be a susceptible target because of its key enzymes, each capable of setting off a cascade of events. For instance, in adult zebrafish exposed to water-soluble fraction of crude oil, Arukwe et al. [13] demonstrated a negative relationship between altered steroidogenic responses and xenobiotic biotransformation processes. The same study showed down-regulation of important steroidogenic enzymes including star and cyp11a1, and up-regulation of hsd3b1 along with increasing contaminant level. Levi et al. [14] showed hepatic expression of star in non-vitellogenic female zebrafish. In the present study we found no change of star mRNA levels except some tendency for down-regulation relative to the control group at exposure levels 5 and 25 ng EE2 L-1 (Figure 2a). Filby et al. [11] also showed down-regulation of star mRNA in adult male fathead minnow exposed to 10 ng EE2 L-1 but no significant changes in exposed females. In vitro cultures of zebrafish ovarian follicles at different maturation stages exposed to human chorionic gonadotropin revealed an increase in expression of star, cyp11a1 and hsd3b1 with mature stages, in contrast with a decreasing pattern without exposure [12]. In this study, the mRNA levels of cyp11a1 and hsd3b1 were relatively lower for all exposure concentrations with lowest relative expression for 25 ng EE2 L-1 at 59% and 45% respectively (Figure 2b). Lower expression levels of star, cyp11a1, and hsd3b1 in fish exposed to EE2, a potent agonist for estrogenic effects, suggest effective regulatory processes in effect. In comparison to three other transcripts, cyp19a1b showed a clear dose-specific up-regulation pattern. Filby et al. [11] that showed upregulation of only cyp19a1a but not cyp19a1b in male fathead minnow exposed to EE2, possibly owing to the pronounced expression of the brain isoform (cyp19a1b) in earlier life-stages. The use of cyp19a1b signal has also been employed in transgenic zebrafish specifically as estrogenic-sensitive screens [44,45]. Because cyp19a1b is a key player in steroid biosynthesis at stage of 20dpfZF, involved in conversion of androgen to estrogen, it qualified as a compelling signature of xenoestrogen-induced effects. Exposure II: cyp19a1b and vtg1 mRNA regulation by four potential xenoestrogens In the present study, modulation of cyp19a1b and vtg1 mRNA expression was used as biomarker for estrogenic effect in the identified responsive ELS of zebrafish (Figure 3). Earlier, in mature adults, estrogen-mediated expression of cyp19a1b has been described among other teleosts, e.g. in Atlantic salmon [46], mangrove killifish [47], medaka [48], and zebrafish [18-20]. Expression of cyp19a1b in developing zebrafish (3-4dpf) was shown to be a sensitive indicator of estrogen-induced effects even earlier than 20dpfZF [49]. Giving further support to our research, Muncke and Eggen [50] described the induction of vtg1 mRNA as a reliable endpoint for 1-5 dpf zebrafish, presenting alternatives for adult animal testing. Figure 3: Exposure II: Xenoestrogen-induced gene expression in 20dpfZF. Fold-change expression ofcyp19a1b andvtg1 in 20dpfZF exposed to three nominal concentrations of EE2 (a), BPA, NP and OP (b) for five days, i.e., over the period 20–25 dpf, in a whole-body assay. Bars represent gene expression fold-change (± SD; n = 3 pools of 15 fish each) normalized to geometric mean ofβ-actinand ef1a and relative to solvent (0.01% methanol) control. The expression levels are in log10 scale. Note expression of both cyp19a1b and vtg1 genes were down-regulated in NP exposure (b). The asterisks represent statistically significant differences, * P< 0.05, ** P< 0.01 and ***P< 0.001 as determined by Welch’s test followed by Games-Howell post hoc test. Citation: Sahoo TP, Oikari A (2013) Use of Early Juvenile Zebrafish Danio Rerio for In-Vivo Assessment of Endocrine Modulation by Xenoestrogens. J Environ Anal Toxicol 4: 202. doi: 10.4172/2161-0525.1000202 Page 7 of 13 Volume 4 • Issue 1 • 1000202 J Environ Anal Toxicol ISSN: 2161-0525 JEAT, an open access journal Our juvenile zebrafish (20 dpf) exposed to EE2 showed inductions of cyp19a1b and vtg1 in a concentration-dependant manner, vtg1 being more distinct than cyp19a1b. Importantly, however, the expression levels of cyp19a1b were increased in all exposure groups relative to control, serving as an excellent benchmark. For comparison, in adult male zebrafish exposed to EE2 for three weeks, induction of cyp19a1b mRNA [51] and protein levels [4] were also evident. Our data with 20dpfZF showed around 13-fold cyp19a1b induction at nominal concentration of 5 ng EE2 L-1 and more at higher concentrations, keeping with range of expression in adults. Even somewhat younger animals than in our study, 17 dpf zebrafish exposed from 0.3 to 30 ng L-1 EE2 showed statistically significant induction in a dose-dependent manner in three days [19,52]. Although significant induction of vtg1 mRNA was observed in very early zebrafish embryos (2 dpf) [50] exposed to 1000 ng EE2 L-1, the same authors also showed significant up-regulation of vtg1 mRNA in zebrafish embryos exposed to 25 ng L-1 for 5 dpf [41]. In 20dpfZF, cyp19a1b was up-regulated in a concentrationdependent manner above 500 µg BPA L-1 (P <0.05) and vtg1 at 1000 µg BPA L-1 (P<0.05). While in adult fish the induction of Vtg has been documented well [4], post-hatched and early juvenile zebrafish exposed to BPA for 5 d induced vtg1 at higher concentrations, 2280 µg BPA L-1 [41]. Our results showed sensitivity of 20dpfZF exposed to 1000 µg BPAL-1 , comparable with adult male zebrafish exposed for three weeks that showed lowest observed effect concentration (LOEC) for vitellogenin protein induction for similar exposure concentration [4]. Thereby, the estrogenic responses of 20dpfZF present a cost-effective life-stage compared to the adult fish as well as further developed metabolic capacity when compared to embryonic stages. Unlike BPA, NP showed clear down-regulation of estrogenic response in 20dpfZF. This is in contrast to significant effects observed in adult [4,53] as well as larval [54] stages of zebrafish and other small bodied fish models, Japanese medaka [55] and fat head minnow [56]. However, other studies this far with zebrafish embryos exposed to NP showed no significant induction of vtg1 mRNA [36,41], indicating low sensitivity to the chemical. Our results confirmed the low sensitivity of juvenile life-stages to NP as shown by Jin et al. [36]. It also contrasts with the significant induction of aromatase by juvenile zebrafish (17 dpf) exposed to NP for 3d [52]. In our study, the mRNA of both cyp19a1b and vtg1 were down-regulated. In essence, the observed discrepancies can be related to the difference in composition of the selected chemical. Previous studies [4,54] with NP included a technical grade of NP (CAS 84852-15-3); a mixture of isomers. In our study, the linear form of 4-n-NP (CAS 104-40-5) that is devoid of branched alkyl side-chains present in most of the isomers, showed slightly reduced estrogenic potency (relative to E2), compared to the higher efficacy of technical grade NP [57]. From metabolite studies of 4-n-NP in fish, biotransformation pathway involving β-oxidation of the alkyl sidechain results in metabolites, lacking the alkyl side-chain, which largely accounts for the lesser estrogenic potential of 4-n-NP in vivo [58]. Among the three alkylphenols in the present study, only OP significantly induced vtg1 in 20dpfZF. Earlier studies with adult zebrafish showed the weak estrogenic potential of OP at similar concentrations as used by us [4]. Thus, considering the differences in sensitivity between Figure 4: Exposure III: Gene induction caused by sediment-borne EE2. Foldchange of cyp19a1b (a) and vtg1 (b) from whole body homogenate of early juvenile zebrafish (23–25 dpf) exposed to sediment spiked with EE2. Fish (n = 3 pools) were sampled at the end of 3 d for Ref-Sed, Water-EE2 (150 ng L-1), Sed 1+Salt and Sed 1 and at the end of 5 d for Sed 2. Bars represent normalized gene expression expressed as fold-change relative to control or reference exposure (Ref-Sed). Values are mean ± SD (15 fish per triplicate pool). Means with different letters are significantly different (P< 0.05) as determined by One-way ANOVA followed by Tukey’s HSD post hoc test. Figure 5: Experimental setup of sediment assay (Exposure III) over 5 day exposure period (a). Free swimming juvenile zebrafish (20-25 dpf) foraging on particulate matter in upper most sediment layer (b). Arrow indicates ingested sediment in accumulated in the gut of fish (body length, 7.8 mm) at the end of exposure (c).