Disruption of endocrine function in H295R cell in vitro and in zebrafish in vivo by naphthenic acids

Journal of Hazardous Materials 299 (2015) 1–9

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Disruption of endocrine function in H295R cell in vitro and in zebrafish in vivo by naphthenic acids Jie Wang a , Xiaofeng Cao a , Jinhua Sun b , Yi Huang a,∗ , Xiaoyan Tang a a State Key Joint Laboratory of Environmental Simulation and Pollution Control, College of Environmental Sciences and Engineering, Peking University, Beijing 100871, China b Chinese Research Academy of Environmental Sciences, Beijing 100012, China

h i g h l i g h t s • • • •

First attempt to study endocrine disruption of NAs via in vivo and in vitro assays. Exposure to NAs inhibited T production, while induced E2 and P4 production in H295R. Key genes for steroidogenesis were significantly altered in NAs exposure groups. CYP19b, ER␣, and VTG were significantly up-regulated in zebrafish exposed to NAs.

a r t i c l e

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Article history: Received 24 April 2015 Received in revised form 1 June 2015 Accepted 2 June 2015 Available online 4 June 2015 Keywords: Naphthenic acids Steroidogenesis Estrogen receptor VTG

a b s t r a c t Oil sands process-affected water (OSPW) have been reported to exhibit endocrine disrupting effects on aquatic organisms. Although the responsible compounds are unknown, naphthenic acids (NAs) have been considered to be implicated. The current study was designed to investigate the endocrine disruption of OSPW extracted NAs (OS-NAs) and commercial NAs (C-NAs) using a combination of in vitro and in vivo assays. The effects of OS-NAs and C-NAs on steroidogenesis were assessed both at hormone levels and expression levels of hormone-related genes in the H295R cells. The transcriptions of biomarker genes involved in endocrine systems in zebrafish larvae were investigated to detect the effects of OS-NAs and C-NAs on endocrine function in vivo. Exposure to OS-NAs and C-NAs significantly increased production of 17␤-estradiol (E2) and progesterone (P4), and decreased production of testosterone (T). Both OSNAs and C-NAs significantly induced the expression of several genes involved in steroidogenesis. The abundances of transcripts of biomarker gene CYP19b, ER␣, and VTG were significantly up-regulated in zebrafish larvae exposed to OS-NAs and C-NAs, which indicated that NAs had negative effects on estrogenresponsive gene transcription in vivo. These results indicated that NAs should be partly responsible for the endocrine disrupting effects of OSPW. © 2015 Elsevier B.V. All rights reserved.

1. Introduction Naphthenic acids (NAs), natural constituents of crude oil, characterized as group of cyclic and acyclic alkyl-substituted carboxylic acids with the general formula Cn H2n+Z O2 , where n is the number of carbons and Z relates to the number of rings [1–3], and may account for as much as 4% of the weight of crude petroleum. As such, NAs are also a major component of oil spills and oil production discharges [4]. Meanwhile, they are frequently found with high concentrations (20–120 mg/l) in oil sands process-affected water (OSPW), which

∗ Corresponding author. E-mail address: [email protected] (Y. Huang). http://dx.doi.org/10.1016/j.jhazmat.2015.06.004 0304-3894/© 2015 Elsevier B.V. All rights reserved.

is generated by extraction of bitumen using the ‘Clark hot water extraction process’, and stored in the mined-out pits with a vast amount in the Athabasca region of Alberta, Canada [5–7]. Actually, toxicities of NAs to organisms were from the research for ecotoxicities of OSPW. Under ‘zero discharge policies’ stipulated by regulatory agencies, oil sand operators cannot release OSPW into the environment [8]. For this reason, manmade tailing ponds are constructed to contain these tailing waters, which then form the end pit lakes. It has been expected that viable ecosystems would develop in these end pit lakes with a biological capability similar to natural lakes in this region [9]. However, considerable investigations into the toxicity of OSPW showed that higher incidences of deformities were observed in yellow perch (Perca flavescens), Japanese medaka (Oryzias latipes), and fathead minnow

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(Pimephales promelas) larvae exposed to fresh OSPW [10,11]. And significant gill and liver histopathological changes were noted in goldfish (Carassius auratus) held in OSPW [12]. Currently, sub-lethal concentrations of OSPW have also shown to cause disturbances to the endocrine function and reproductive system. Exposure to OSPW decreased concentrations of testosterone (T) and estradiol (E2) in plasma of yellow perch (P. flavescens) [13], goldfish (C. auratus) [14], and impaired the reproductive capacity of fathead minnows (P. promelas) as exemplified by decreased fecundity, altered synthesis of sex steroid hormones, and less pronounced secondary sex characteristics of male and female minnows [15]. Abundances of transcripts of regulatory genes in all tissues of the brain–gonad–liver axis were significantly different in fathead minnows (P. promelas) exposed to OSPW [16]. Meanwhile, in vitro studies have demonstrated that OSPW could cause endocrine disrupting effects on sex synthesis and receptor signaling [17,18], and induce multiple transcriptional changes in trout hepatocytes including vitellogenin (VTG) and estrogen receptor (ER) [19,20]. Although the components of OSPW responsible for the endocrine disrupting effects are unknown, NAs have been considered as the primary candidate contaminants [14,16,21]. In vitro study has previously demonstrated that NAs are weak estrogenic receptor agonists and potent androgen receptor antagonists [22]; however, exposure to NAs did not have a statistically significant effect on VTG production in male stickleback in vivo [23]. The inconsistent results between in vitro and vivo assays indicated that there are still major knowledge gaps in understanding of the capacity and mechanisms of endocrine disrupting effects of NAs. The growing concern over endocrine disrupting chemicals (EDCs) has emphasized the development of screening assays that can address the potential effects on human health. The human adrenocoritical cell line, H295R, shows characteristics of an undifferentiated foetal adrenal cortex and is capable of full steroidogenesis, and it has been employed as an in vitro model for studying endocrine disruptors [24–26]. Meanwhile, in vivo studies are needed to verify the results of in vitro studies. The zebrafish embryo/larvae has become a systematic, sensitive and easily operated mode animal for identify endocrine disruption [27]. Several zebrafish genes, such as CYP19a, CYP19b, ER␣, and VTG are wellknown biomarker genes responsive to environmental endocrine disruptors [28–32]. Therefore, a combined in vitro and in vivo approach is an efficient way to gain a complete understanding of the properties of endocrine disruptors. To investigate the endocrine disrupting effects and the underlying mechanisms of NAs, a combination of in vitro and in vivo assays was employed in the current study. In the in vitro H295R assay, production of the steroid hormones testosterone (T), 17␤estradiol (E2) and progesterone (P4), and abundances of transcripts of the genes involved in steroidogenic pathways were examined. In the in vivo zebrafish (Danio rerio) assay, the expressions of estrogen-responsive biomarker genes, P450 aromatase (CYP19a and CYP19b), estrogen receptors (ER␣, ER␤1, and ER␤2), and vitellogenins (VTG) were investigated in the early life stage of zebrafish exposed to NAs. This study will help better understand the effects of chronic exposure to OSPW on endocrine disruption potentials, and identification of compounds most closely associated with adverse effects of OSPW.

2. Materials and methods 2.1. Exposure NAs preparation Oil sand samples (OS) were a gift from SPT Energy Group Inc. (China) in Daqing oil exploring area. The Oil Sand NAs extraction (OS-NAs) was modified according to Holowenko et al. and Gagne

et al. [5,19]. An industrial bitumen-extraction scheme based on the method of Clark was performed in the laboratory [19]. A sample of oil sand was digested in the presence of 0.1 mol/l NaOH (solid:liquid ratio = 1:2), and heated up to around 70 ◦ C. The mixture was aerated for 10 min to allow the bitumen to partition to the surface, and then discarded the bitumen flocculation. The resulting slurry was cleaned of solid by centrifugation, and the NAs were existed in the supernatant, which was identified as OSPW. The water phase was acidified with HCl to pH < 2 and extracted with dichloromethane (DCM). The extracts were dried with Na2 SO4 and filtered through a glass filter funnel, and then evaporated to dryness under nitrogen gas flow. The residues were dissolved in dimethylsulfoxide (DMSO) to concentration of 1 mg/ml as exposure stock solution. One commercial petroleum derived NAs (C-NAs) were purchased from Sigma–Aldrich (No. 70,340–250 ml), and dissolved in DMSO to concentration of 1 mg/ml as exposure stock solution.

2.2. H295R cell culture and exposure Human adrenocortical carcinoma H295R cells purchased from ATCC (CRL-2128, ATCC, USA) were maintained in DMEM/F12 medium containing HEPES buffer, l-glutamine, pyridoxine HCl (Invitrogen, USA), supplemented with 1% ITS + Premix (BD Bioscience, USA) and 2.5% NuSerum (BD Bioscience, USA). The cultures were kept at 37 ◦ C with 5% CO2 , and the medium was changed three times a week. For chemical exposure, H295R cells were seeded into 24-well plates (Corning, USA) at a concentration of 3 × 105 cells/ml in 1 ml of medium per well. After 24 h culture, the medium was refreshed, and the cells were exposed to OS-NAs (0, 0.01, 0.1, 1, and 10 mg/l) or C-NAs (0, 0.01, 0.1, 1, and 10 mg/l) for 48 h. Dimethyl sulphoxide (DMSO) was used as the carrier solvent and did not exceed 0.01% (v/v). The culture medium and remaining cells were used for hormone measurements and gene transcription analysis, respectively. All treatments were employed in triplicates. Cytotoxicity of OS-NAs and C-NAs was assessed in 96-well plates using the MTT, and no significant differences were observed between NAs exposure groups and the controls, suggesting that there was no significant effect on the cell viability.

2.3. Zebrafish maintenance and exposure Zebrafish (wild type, AB stain) were reared in the Zebrafish Research Lab, Peking University, and maintained under routine approved animal welfare protocols with standard laboratory conditions (at 28.5 ◦ C and 14 h light–10 h dark cycle). Adult fish were fed three times daily with live brine shrimp nauplii, Artemina sp. Five sexually mature males and ten sexually mature females were placed in a breading tank in the afternoon the day before the experiment and embryos were spawn synchronously the next morning. Fertilized and healthy embryos were selected by microscopic observation. Embryos were placed in 6-well plates (ten embryos/well) at 1 h post fertilization (hpf) containing 2 ml NAs test solution/well. Different concentrations of NAs were chose according to the preliminary study (0, 0.05, 0.5, 2.5 mg/l). Fifty percent of the volume was replaced daily with fresh test solutions. A solvent and a non-solvent control were included to confirm that DMSO concentration used had no effect on zebrafish development. All treatments were performed at 28 ◦ C with a 14/10 h light/dark cycle. Any dead embryos or larvae were removed daily. Every exposure (168 hpf) was carried out in triplicate (3 plates). After exposure, fish larvae were collected by sieve, transferred to 2.0 ml cryogenic vials, and placed on ice for rapid euthanasia. Excess water was removed, and samples were stored at −80 ◦ C until further analysis.

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2.4. Hormone measurement The concentrations of T, E2, and P4 were measured by competitive enzyme-linked immunosorbent assay (ELISA), with commercially available kits (T, cat# 582,701, E2, cat# 582,251, P4, cat# 582,601, Cayman Chemical, USA). Briefly, hormones were extracted from the H295R culture media using diethyl ether or methylene chloride, and the extracts were subsequently employed for ELISA following the manufacturer’s instructions. Absorbance of the extracted samples and the hormone standards were measured using Synergy H1M (BioTek, USA) at a wavelength of 415 nm. All groups were measured in triplicate, and the relative fold-changes were determined.

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final round of amplification to differentiate between desired PCR products and primer-dimers or DNA contaminants. 2.6. Statistical analysis Statistical analysis was conducted by SPSS 16.0, and all data were expressed as mean ± SEM. The homogeneity of variance was assessed using Levene’s test. Statistical differences were evaluated by one-way ANOVA followed by posthoc Turkey’s test. Differences were considered statistically significant at p < 0.05. 3. Results 3.1. Concentrations of hormones

2.5. Quantitative real-time PCR (RT-PCR) RT-PCR was used to assess differences in abundances of transcripts of target genes among treatment groups. For in vitro assay, the modulation of 8 major genes involved in the synthesis of steroid hormones (HMGR, StAR, CY11a, CY17, 3␤HSD2, 17␤HSD1, 17␤HSD4, and CYP19a) was investigated after OS-NAs and C-NAs exposure. For in vivo assay, abundances of expressions of P450 aromatase (CYP19a and CYP19b), estrogen receptors (ER␣, ER␤1, and ER␤2), and vitellogenins (VTG) were quantified as markers of endocrine disruption. Sequences of primers used for quantification of transcripts are listed in Table 1. The fold changes in the target gene expression level were normalized to the ␤-actin contents. The ␤-actin was selected as the housekeeping gene because the mRNA expression of this gene was not affected after NAs exposure (data not shown). The relative gene expression was calculated by the 2−Ct methods [33]. For in vitro assay, total RNA was isolated from H295R cells harvested from each single well using TRIzol reagent (Invitrogen, USA) following the manufacturer’s protocol. RNA was quantified by use of a NanoDrop ND-1000 spectrophotometer (Thermo Scientific, USA) after DNase treatment (Qiagen, UK). The first strand cDNA was synthesized from 1 ␮g of total RNA using a PrimerScript RT reagent Kit (Takara, Japan) according to the manufacturer’s introductions. The cDNA samples were stored at −80 ◦ C until further analysis. RT-PCR was performed on an ABI 7500 Real-Time PCR system in 96-well PCR plates (Applied Biosystems, USA). A PCR reaction mixture for one reaction contained 12.5 ␮l 2X SybrGreen qPCR Master Mix (ABI, USA), 1 ␮l forward and reverse primers, 9.5 ␮l RNase free water, and 2 ␮l the first stand cDNA. The PCR reaction mix was denatured at 95 ◦ C for 10 min before the first PCR cycle. The thermal cycle profile was denaturizing for 15 s at 95 ◦ C, annealing for 30 s at 60 ◦ C, and extension for 30 s at 72 ◦ C for a total 40 PCR cycles. Melting curve analyses were performed after the final round of amplification to differentiate between desired PCR products and primer-dimers or DNA contaminants. For in vivo assay, total RNA was extracted from 20 larvae using TRIzol reagent (Invitrogen, USA) following the instructions of the manufacturer. Purified RNA was quantified by use of a NanoDrop ND-1000 spectrophotometer (Thermo Scientific, USA). The first strand cDNA was synthesized from 1 ␮g of total RNA using a PrimerScript RT reagent Kit (Takara, Japan) according to the manufacturer’s introductions. The cDNA samples were stored at −80 ◦ C until further analysis. RT-PCR was performed on an ABI 7500 Real-Time PCR system in 96-well PCR plates (Applied Biosystems, USA). A PCR reaction mixture for one reaction contained 12.5 ␮l 2X SybrGreen qPCR Master Mix (ABI, USA), 1 ␮l forward and reverse primers, 9.5 ␮l RNase free water, and 2 ␮l the first stand cDNA. The PCR reaction mix was denatured at 95 ◦ C for 10 min before the first PCR cycle. The thermal cycle profile was denaturizing for 15 s at 95 ◦ C and annealing and extension for 1 min at 60 ◦ C for a total 40 PCR cycles. Melting curve analyses were performed after the

Generally, both OS-NAs and C-NAs significantly decreased production of T, and increased production of E2 and P4 after 48 h exposure (Fig. 1). There was a significant dose-response relationship in OS-NAs exposure, with E2 production increasing with increasing concentration of OS-NAs (r2 = 0.991). A significant increase in E2 and P4 productions (1.67 and 1.38 folds, respectively) was observed in cells exposed to 10 mg/l OS-NAs, while a significant decrease in T production was found in cells exposed to the highest concentration of OS-NAs, with 0.51 fold change. For C-NAs exposure, similar results were observed. C-NAs significantly induced the production of E2 at 1 and 10 mg/l concentrations, with 2.14 and 1.61 folds change, respectively. Exposure to 10 mg/l C-NAs significantly increased production of P4 (1.31 fold) and decreased production of T (0.38 fold). 3.2. Gene expression responses in H295R cells The expression levels of 8 steroidogenic genes in H295R cells exposed to OS-NAs and C-NAs were determined. Significant differences were observed among the OS-NAs exposure groups and the controls for most of the genes measured (Fig. 2). The abundances of transcripts of StAR and 3␤HSD2 were significantly induced in H295R cells exposed to 10 mg/l OS-NAs, with 19.39 and 27.67 fold, respectively. While the relative expression of HMGR, CYP11a, CYP17, 17␤HSD1, and CYP19a, to a less extent, were significantly induced by 6.02, 2.98, 4.22, 4.55, and 4.38 fold, respectively, at 10 mg/l OS-NAs exposure groups. Exposure to OS-NAs increased the expression of 17␤HSD4, but the changes were not significantly different. The expression pattern for the 8 steroidogenic genes was also altered after C-NAs exposure (Fig. 3). Expression levels of StAR, 3␤HSD2, and CYP19a were significantly increased at 10 mg/l C-NAs exposure groups, with 16.15, 25.58, and 23.59 fold, respectively. For HMGR, CYP11a, CYP17, and 17␤HSD1, the relative responses were significantly induced to a less extent (6.99, 2.46, 3.28, and 3.56 fold, respectively). There was no significant difference in 17␤HSD4 gene expression between C-NAs exposure groups and the controls. 3.3. Gene expression responses in zebrafish Different responses of marker genes related to endocrine disruption were observed in zebrafish larvae exposed to OS-NAs and C-NAs. Abundances of transcript of ER␣ were significantly induced by high concentrations of OS-NAs, with 10.23 fold greater for 0.50 mg/l and 5.91 fold change for 2.50 mg/l. No significant change was observed in ER␤1 gene expression in larvae exposed to OS-NAs compared to the controls, while only 0.50 mg/l OS-NAs significantly induced the expression of ER␤2 (Fig. 4). There was no significant difference in the abundance of transcripts of CYP19a among exposure groups. However, the abundance of transcripts of CYP19b were significantly induced in larvae exposure to three concentration OSNAs, with 23.85, 33.86, and 39.06 fold changes for 0.50 0.05, and

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Fig. 1. Concentrations of 17␤-Estradiol (E2), progesterone (P4), and testosterone (T) in the culture medium of H295R cells exposed to OS-NAs and C-NAs exposure groups (n = 3). Asterisks represent a significant difference compared with the control (p < 0.05).

Fig. 2. The effects of OS-NAs on steroidogenic gene expression in H295R cells after 48 h exposure (n = 3). The mRNA expression levels are compared to the controls. Asterisks represent a significant difference compared with the control (p < 0.05).

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Table 1 Primer sequences of the genes tested in the H295R cell and zebrafish study. Sequence of the primers (5 #4 3#)

Gene name Forward

Reverse

H295R

␤-actin [34] 3␤HSD2[34] 17␤HSD1[34] 17␤HSD4[34] CYP11a [34] CYP17 [35] CYP19a [35] HMGR [35] StAR [35]

CACTCTTCCAGCCTTCCTTCC TGCCAGTCTTCATCTACACCAG CTCCCTCTGACCAGCAACC TGCGGGATCACGGATGACTC GAGATGGCACGCAACCTGAAG AGCCGCACACCAACTATCAG AGGTGCTATTGGTCATCTGCTC TGCTTGCCGAGCCTAATGAAAG GTCCCACCCTGCCTCTGAAG

AGGTCTTTGCGGATGTCCAC TTCCAGAGGCTCTTCTTCGTG TGTGTCTCCCACGCAATCTC GCCACCATTCTCCTCACAACTC CTTAGTGTCTCCTTGATGCTGGC TCACCGATGCTGGAGTCAAC TGGTGGAATCGGGTCTTTATGG AGAGCGTTCGTGGGTCCATC CATACTCTAAACACGAACCCCACC

Zebrafish

␤-actin [36] ER␣ [37] ER␤1 [37] ER␤2 [37] CYP19a [38] CYP19b [38] VTG [36]

ACCCACACCGTGCCCATCTA GGTCCAGTGTGGTGTCCTCT TTGTGTTCTCCAGCATGAGC TAGTGGGACTTGGACCGAAC AGATGTCGAGTTAAAGATCCTGCA GACACTCTCTCCATCAGTCTGTTCTT AACGAACAGCGAGAAAGAGATTG

CGGACAATTTCTCTTTCGGCTG AGAAAGCTTTGCATCCCTCA CCACATATGGGGAAGGAATG TTCACACGACCACACTCCAT CGACCGGGTGAAAACGTAGA CATTCAGTTTCTGCAAGTCAGCA GATGGGAACAGCGACAGGA

Fig. 3. The effects of C-NAs on steroidogenic gene expression in H295R cells after 48 h exposure (n = 3). The mRNA expression levels are compared to the controls. Asterisks represent a significant difference compared with the control (p < 0.05).

2.50 mg/l, respectively (Fig. 4). For VTG gene, exposure to higher concentrations of OS-NAs (0.50 and 2.50 mg/l) significantly induced the expression of VTG gene in zebrafish larvae, with 3.84 and 2.02 fold greater compared to the controls, respectively. Similar alteration was observed in zebrafish larvae exposed to C-NAs (Fig. 5). ER␣ gene expressions were significantly greater by 6.82 and 3.88 fold, respectively, in larvae exposed 0.05 and 0.50 mg/l concentrations of C-NAs. And significant change in ER␤2 gene expression could be observed in larvae exposed to 0.05 mg/l C-NAs. There was no significant different in abundances of transcript of ER␤2 in larvae between C-NAs exposure groups and the controls. For genes encoded P450 aromatase, no significant change was found in CYP19a gene expression in larvae exposed to C-NAs, while the abundance of transcripts of CYP19b were significantly greater compared to the controls, with 6.13, 9.64, and 4.10 fold changes for 0.05, 0.50, and 2.50 mg/l, respectively. And also, significant changes of VTG gene were found in larvae exposed to 0.50 and 2.50 mg/l concentrations of C-NAs (2.23 and 3.10 fold, respectively).

4. Discussion The endocrine disruption of OSPW has been well documented in vitro [17–20] and in vivo [9,13–16,21]; however, the compounds responsible for the activities in OSPW were not assigned. Although, NAs has been considered as the primary toxic compounds in OSPW, few studies have directly performed the endocrine disrupting function of NAs. The objective of the current study was to evaluate the role of NAs in disrupting the endocrine system. A combination of in vitro and in vivo assays was used to explore the endocrine disrupting function of NAs. NAs might disrupt the endocrine function in H295R cells via steroidogenic modulation. H295R cells have been used for rapid screening of potential disruption of steroidogenic pathways as well as determining mechanisms of action of toxicants [39–42]. And determination of steroid hormones released into the medium has been suggested to be the most integrative, functional endpoint and has been established previously in H295R cells [25,26,43]. In the

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Fig. 4. The effects of OS-NAs on marker gene expression of endocrine disruption in zebrafish larvae. Quantification was performed on three independent experiments, each experiment was employed in triplicates. The mRNA expression levels are compared to the controls. Asterisks represent a significant difference compared with the control (p < 0.05).

Fig. 5. The effects of C-NAs on marker gene expression of endocrine disruption in zebrafish larvae. Quantification was performed on three independent experiments, each experiment was employed in triplicates. The mRNA expression levels are compared to the controls. Asterisks represent a significant difference compared with the control (p < 0.05).

current study, significantly inhibited production of T and induced production of E2 and P4 were observed in H295R cells after exposure to OS-NAs and C-NAs. Previous study has concluded that untreated OSPW significantly decreased T and increased E2 concentrations in H295R cells medium [17]. And the decrease in T production in H295R cells is also consistent with lesser concentration of T in blood plasma in vivo [9]. HMGR, StAR, and CYP11a

are rete-limiting steps in the conversion of cholesterol to steroid hormones. Gene expression levels of HMGR, StAR, and CYP11a were significantly induced in H295R cells after exposure to OS-NAs (Fig. 2) and C-NAs (Fig. 3), which might be the indirect reason of increased production of P4. The gene expression of 3␤HSD2 was also induced in NAs exposure groups, which indicated that OS-NAs and C-NAs could direct increase P4 production by up-regulating

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3␤HSD2 gene expression. 3␤HSD2 is an essential enzyme in the biosynthesis of mineralocorticoids, glucocorticoids and sex hormones [26]. Alteration of 3␤HSD2 expression implied that both OS-NAs and C-NAs might have influences in other steroidogenic pathways, including sex hormones synthesis. CYP17 is essential for the conversion of P4 to T. The fact that T production was not affected combined with the up-regulation of CYP17 transcription might implied that production of steroids is a complex process with multiple sensitive control points, and simple statistical correlations between gene expression and hormone production should not be expected [24]. Inhibited production of T and enhanced production of E2 was observed in H295R cells exposed to OS-NAs and C-NAs, and this was accompanied by up-regulation of gene expression of P450 aromatase (CYP19a). CYP19a catalyzes the final and rate-limiting step in conversion of androgen to estrogen, and thus, the up-regulation of CYP19a gene expression could explain the alteration of production of T and E2 in the presence of OS-NAs and C-NAs. This result is consistent with He et al.’s study that the aromatase mRNA abundance and enzyme activity were significant greater in cells exposed to OSPW [17]. Alteration of sex hormones production and genes transcription in H295R cells suggested that the endocrine disrupting function caused by OSPW in vitro should be partly attributed to the NAs in OSPW. Zebrafish can serve as a suitable model for assessing xenogenous chemicals as disruptors of the endocrine system. However, according to the 3Rs principle, animal tests should be reduced, refined and, if possible, replaced. In this context, zebrafish embryos/larvae are suitable alternatives since they are considered less susceptible to distress and pain than adults [27]. More important, zebrafish embryos/larvae express endocrine markers such as aromatase and vitellogenin that are affected by endocrine compounds, allowing them to be tested by monitoring the transcriptome [31,32]. Therefore, zebrafish embryos/larvae were employed to investigate the endocrine disruption of NAs in vivo in this study. Our results showed that both OS-NAs and C-NAs could affect the genes expression involved in endocrine function system in the early life stage of zebrafish. In the current study, OS-NAs and CNAs altered gene expression patterns of estrogen receptors (ERs). There are three types of ERs, characterized as ER␣, ER␤1, and ER␤2 in zebrafish [44,45], and different EDCs have been proved to affect the ERs gene expression in zebrafish. The natural ligand 17␤-estradiol can bind all three ERs with similar affinities [44]. The abundances of transcripts of ER␣ were significantly induced in zebrafish exposed to other synthetic ligands, such as diethylstilbestrol, 4-nonylphemol, PFOS, and PFOA [28–30,46]. In this study, we analyzed expression patterns of the three ER genes. Significantly high expressions of ER␣ were detected in the OS-NAs and C-NAs exposure groups. No significant difference was observed in ER␤1 and ER␤2 gene expressions in NAs exposure groups and the controls, except for ER␤2 in 0.50 mg/l OS-NAs group and 0.05 mg/l C-NAs group (Figs. 4 and 5). It has been suggested that OSPW could cause estrogenic effect through receptor mediated pathways in T47D-kbluc cells in vitro [18]. And the abundances of transcripts of estrogen-related receptor genes were significantly altered in Chironomus dilutes larvae and Fathead minnows (P. promelas) exposed to OSPW [16,21]. According to these previous results and the current results, it is clearly that NAs could be considered as the candidate components in OSPW as endocrine disruptors via receptor modulation. Although no significant change was found in the transcriptions of CYP19a in zebrafish larvae exposed to OS-NAs and C-NAs, upregulation of CYP19b gene was observed in the exposure groups (Figs. 4 and 5), which could enhance the activity of P450 aromatase, and promote the conversion of androgens to estrogens. This result is consistent with the in vitro assay in the current study that CYP19a gene expression was significantly induced in H295R

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cells in the exposure groups. Combining in vitro and in vivo assays about steroidogenic disruption of NAs, it can be concluded that NAs cause endocrine disrupting effects via steroidogenic modulation, which might be the possible reason for lesser concentrations of T in blood plasma in previous OSPW studies [13–15]. Transcription of VTG gene has been developed as a biomarker for estrogenic pollution [47–49]. VTG gene expression could be significantly altered in zebrafish larvae by endocrine disrupting chemicals, such as pesticides, polychlorinated biphenyls (PCBs), bisphenol A [27,52]. In current study, high concentrations of OS-NAs and C-NAs significantly induced the abundances of transcripts of VTG in larvae, with 3.84 fold greater for 0.50 mg/l OS-NAs and 3.10 fold greater for 2.5 mg/l C-NAs (Figs. 4 and 5). The up-regulation of VTG gene indicated that both OS-NAs and C-NAs are estrogenic to zebrafish larvae. The plausible explanation of the up-regulation might be that NAs work as direct-acting ER agonists, and then stimulate the transcription of VTG gene. This is consistent with the observation that ER␣ was transcriptionally significantly up-regulated in zebrafish larvae exposed to OS-NAs and C-NAs. In vitro assays could provide screening and classification of potential endocrine disrupting of chemicals; however, it is a still challenge to validate results from in vitro assays to predict these responses in vivo for the differences in the complexity of the two systems (e.g., cells versus tissues) and the differences in species [50,51]. In the current study, the transcriptions of P450 aromatase in H295R cell line and zebrafish larvae were both up-regulated in NAs treatments. However, it is still quite difficult to rely on in vitro data alone to predict the endocrine disrupting properties of xenobiotics, due to the different mechanisms and the complexity of endocrine function in animals [40,51]. Therefore, it is necessary to combine the effects of chemicals in vitro and in vivo. Further study in order to evaluate the genes and hormones involved in endocrine function in adult zebrafish will be employed. NAs has been considered to be primarily responsible for the adverse effects of OSPW in aquatic organism, such as acutely lethal effects, oxidative stress, and endocrine disruption. However, it is still a confusion that whether commercial NAs are an adequate model to study OS-NAs toxicity in complex organisms. Several commercial NAs have been used to assess the toxicity in vitro and in vivo. NAs purchased from Sigma were employed to evaluate the development and incidence of deformities in amphibian larvae [53]. And Zhang et al. has assessed the toxicity of a technical mixture of NAs purchased form Sigma using a microbial genome wide live cell reporter array system. Up-regulation of genes in the pentose phosphate pathway and down-regulation of the ATP-binding cassette transporter complex was observed [54]. The androgen receptor antagonist potencies with NAs purchased from Fluka and Acros were also observed in Thomas et al.’s study [22]. However, West et al. has declared that the unknown impurities in C-NAs mixtures may limit the usefulness of C-NAs for toxicity assays [55]. And also, Garcia-Garcia et al. compared the effects of C-NAs and OSPW NAs in the pro-inflammatory genes expressions in mice, and found that C-NAs may not necessarily reflect the possible effects induced in organisms [56]. In the current study, we studied the endocrine disruption in vitro and in vivo assays with OS-NAs and C-NAs. The alteration of sex hormones production and the transcription pattern in H295R cells exposed to C-NAs were similar with the changes in OS-NAs exposure groups. And also, similar changes of genes expression involved in endocrine system were observed in zebrafish larvae exposed to OS-NAs and C-NAs. It seems that there was no toxic difference between OS-NAs and C-NAs. However, significant differences were observed in the relative gene expression of CYP19 in H295R cells and in zebrafish larvae. It is suggested that there is a wider range of components in OS-NAs than C-NAs. Recently, the term “acid extractable fraction” has been introduced as a more accurate terminology as normal procedures to extract

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NAs from environmental samples also include non-classical NAs in addition to a complex, not fully characterized mixture [57,58]. The acid extractable fraction may consist not only of naphthenic acids, with other compounds such as PAHs and alkyl phenols [59,60]. The toxic differences may be caused by these non-NAs compounds, and this inference will be studies in the future. C-NAs used in the current study is therefore not regarded as representative for the OS-NAs; however, it does nonetheless provide an assessment on the toxicity of oil sand pollution. Because the alteration of gene expression patterns were quite the same.

5. Conclusion In summary, our study confirmed that NAs is able to have multiple effects on the endocrine system through interfering with hormone receptor ERs, altering steroid hormone synthesis and the steroidogenic genes expression in H295R cell line, and disrupting the gene transcription levels involved endocrine system in zebrafish. Both OS-NAs and C-NAs affects the hormone production by changing the expression of several key steroidogenic genes in H295R cell line in vitro. And in the vivo experiment, biomarker genes related to endocrine systems, such as ER␣, CYP19b, and VTG were significantly induced in OS-NAs and C-NAs exposure. Combining in vitro and in vivo assays, it is suggested that NAs could be considered as the primary candidate components in OSPW related to the endocrine disrupting function. The similar effects in OS-NAs and C-NAs exposure groups implied that C-NAs could be used as model chemicals to study estrogenic toxicity of NAs from OSPW.

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