Chronic exposure to environmental concentrations of phenanthrene impairs zebrafish reproduction

Ecotoxicology and Environmental Safety 182 (2019) 109376

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Chronic exposure to environmental concentrations of phenanthrene impairs zebrafish reproduction

T

Xiandong Peng, Xiaoxi Sun, Min Yu, Wei Fu, Hua Chen, Jiazhou Chen∗ Obstetrics & Gynecology Hospital of Fudan University, Shanghai JiAi Genetics & IVF Institute, Shanghai, China

A R T I C LE I N FO

A B S T R A C T

Keywords: Phenanthrene (PHE) Endocrine disruption Reproduction Sex hormones Zebrafish

Phenanthrene (PHE) is a tricyclic polycyclic aromatic hydrocarbon which distributed extensively in the aquatic environment. However, the knowledge about its impact on fish reproduction is still limited, particularly under a chronic exposure regime. In this study, we exposed zebrafish (Danio rerio) embryos to environmentally relevant concentrations (0.2, 1.0, and 5.0 μg/L) of PHE for 4 months and assessed the impact on reproduction. The results demonstrated that egg production was decreased in fish exposed to PHE, with a significant reduction at 5.0 μg/L. The exposure significantly decreased the circulating concentrations of estradiol (E2) and testosterone (T) in female fish or E2 in male fish. In addition, plasma vitellogenin levels were significantly inhibited after PHE exposure in female fish. The transcription of hypothalamic-pituitary-gonadal (HPG) axis related genes (GnRH2, FSHβ, LHβ, 17β-HSD, CYP11A1, and CYP19a) were significantly altered in a sex-specific manner. In addition, embryos derived from exposed parents exhibited increased malformation and decreased hatching success in the F1 generation. Taken together, these results demonstrate that chronic exposure to environmentally relevant concentration of PHE could cause adverse effects on reproduction and impair the development of offspring, ultimately leading to fish population decline in aquatic environment.

1. Introduction

water from the Jiulong River Estuary and Western Xiamen Sea (Maskaoui et al., 2002). The maximum reported concentrations in water were 1460 μg/L for PHE in a fishing settlement near a crude oil exploration area in Nigeria (Anyakora et al., 2005). Previous studies indicates that several PAHs are endocrine disruptors and capable of altering hormonal equilibrium, consequently causing reproductive and developmental toxicity on diverse organisms (Archibong et al., 2012; Borman et al., 2000; Nicolas, 1999; Zenzes, 2000). For example, benzo(a)pyrene (BaP) has consistently been shown to have a negative impact on fish reproduction such as reduced cumulative egg production, impaired ovarian growth and lower circulating levels of testosterone (T) and 17β-estradiol (E2) (Hoffmann and Oris, 2006; Thomas, 1990; Zheng et al., 2006). Previous in vitro study demonstrated that PHE inhibits androstenedione (A) and E2 synthesis through blocking the activity of a rate-limiting enzyme for conversion of C21 to C19 steroids (Monteiro et al., 2000b). In female flounder (Platichthys flesus), exposure to PHE via diet (0.5, 2.5 or 12.5 nmoL/g food) for 12 weeks shows decreased plasma E2 concomitant with increased levels of plasma 17α-hydroxyprogesterone (17α-OHP) (Monteiro et al., 2000a). Similarly, in male Sebastiscus marmoratus, chronic exposure to PHE at environmentally relevant concentrations

Polycyclic aromatic hydrocarbons (PAHs) are a group of ubiquitous environmental contaminants that are primarily produced from incomplete combustion or pyrolysis of organic material such as wood, coal, fuel, tobacco, and meat (Bostrom et al., 2002). Recently, economic growth, advance of urbanization and increased energy consumption have resulted in growing PAHs released into the environment. Owing to their low solubility, low vapor pressures, and lipophilic in nature, PAHs are resistant to biodegradation and can readily bioaccumulate in human tissues, such as milk, placenta, as well as circulate in umbilical cord blood (Yu et al., 2011b). Among these PAHs, phenanthrene (PHE), a three-ring PAH included in the 16 US EPA priority PAHs, is one of the most abundant PAHs in the aquatic environment as a result of human activities. PHE inputs into the aquatic environment mainly through water runoff, atmospheric deposition and oil spills (Incardona et al., 2005; Lima et al., 2003). It is reported that the concentration of PHE in the surface water in Tianjin and Daya Bay ranged from 531 to 1420 ng/ L and 153–1445 ng/L, with a mean concentration of 887 ng/L and 435 ng/L, respectively (Cao et al., 2005; Zhou and Maskaoui, 2003). Greater concentration of PHE (2.44–26.1 μg/L) was reported in the pore

∗ Corresponding author. Obstetrics & Gynecology Hospital of Fudan University, Shanghai JiAi Genetics & IVF Institute, NO.588, Fangxie Road, Huangpu District, Shanghai, China. E-mail address: [email protected] (J. Chen).

https://doi.org/10.1016/j.ecoenv.2019.109376 Received 22 April 2019; Received in revised form 18 June 2019; Accepted 20 June 2019 0147-6513/ © 2019 Elsevier Inc. All rights reserved.

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shows reduced E2 levels and γ-GTP activity in the testes (Sun et al., 2011). Steroid hormones are derived from cholesterol through multi-steps reactions catalyzed by a series of enzymes, and exert their effects by binding to specific intracellular receptors that acts as ligand-inducible transcription factors and regulating the expression of target genes (Taghizadeh et al., 2013). In fish and other vertebrates, the hypothalamic-pituitary-gonadal (HPG) axis and steroid hormones play critical roles in multiple reproductive processes such as gametogenesis and reproductive maturations (Sofikitis et al., 2008; Sower et al., 2009). Previous studies conducted in fish have demonstrated that the HPG axis is a powerful tool for assessing endocrine-disrupting effects and exploring the potential mechanisms in toxicological research (Villeneuve et al., 2008; Zhang et al., 2008). Disruption at any steps in the process of steroid biosynthesis may have adverse effects on endocrine system and reproduction. For example, growing evidence demonstrated that fish exposed to chemicals disturbing steroidogenesis pathway exerted adverse reproductive effects (Ankley et al., 2005, 2007; Skolness et al., 2013; Villeneuve et al., 2008). Recently, PHE has attracted considerable attention because of its increasing environmental concentrations. Moreover, its unique physical and chemical properties, such as low molecular weight and relatively low Kow facilitate it to be dissolved in water and adsorbed to particles. As a consequence, PHE is more likely to be bioaccumulated in aquatic wildlife via water and sediment sources (Loughery et al., 2018). However, the ecotoxicological effect of PHE on fish is still limited, particularly under a chronic exposure regime. In the present study, we exposed zebrafish embryos to environmental concentration of PHE (0.2, 1.0, and 5.0 μg/L) and investigated the impact of long-term PHE exposure on endocrine status and reproduction. We employed zebrafish because this species shares extensive conserved syntenic fragments with human and has been widely used as an appropriate model for the study of chemical toxicity (Hill et al., 2005). Several toxicological endpoints such as developmental toxicity, egg production, plasma levels of VTG as well as sex hormones (T and E2), and transcriptional profiles of genes related to steroidogenesis were examined in this study.

samples ranged from 86.2% to 102%. The limit of detection (LOD) and the limit of quantitation (LOQ) of the HPLC analysis method were 0.1 μg/L and 0.33 μg/L, respectively. 2.2. Fish maintenance and experimental design Male and female adult zebrafish (AB-type 3–4 mo old) were maintained and raised in a semi-static system that received charcoal-filtered tap water at a constant temperature (28 ± 1 °C), with a photoperiod 14:10 (light:dark). The adult fish were fed newly hatched brine shrimp (Artemia nauplii) and a commercial food pellet (Trea, Germany) once daily in a quantity that was consumed within 5 min. Zebrafish embryos were obtained from spawning adults in a ratio of 1:2 (male: female) in tanks overnight. Embryos that had developed normally and reached the blastula stage (2 h post-fertilization, hpf) were selected for subsequent experiments. The embryos were randomly distributed into separate glass beakers containing 400 mL of exposure solution (0, 0.2, 1.0, and 5.0 μg/L PHE) containing 0.2 mM Ca (NO3)2, 0.13 mM MgSO4, 19.3 mM NaCl, 0.23 mM KCl and 1.67 mM HEPES. The selected exposure concentrations were based on the results of a previous experiment to detect the highest concentration of PHE that did not cause acute developmental toxicity (Huang et al., 2013). There were three replicates for each exposure concentration, with each beaker containing 100 embryos. At 21 days post-fertilization (dpf), the larvae were transferred into 20-L tanks. At 60 dpf, the fish were transferred into 30-L tanks containing 20 L of PHE solution with the same concentration. At approximately 100 dpf, 9 males and 9 females were randomly selected from each control and exposure tank. The fish were paired and eggs in each tank were collected daily. The total number of eggs produced in three consecutive weeks were summed and recorded. During the experimental period, half of the water in each tank was replaced with fresh water daily and adjusted to the appropriate PHE concentration. After the exposure, fish were euthanized in 2-phenoxyethanol (SigmaAldrich, St. Louis, MO, USA) and total weight and snout-to-vent length was recorded for each fish. Indices including condition factor (K; total wt × 100/total length3), brain-somatic index (BSI; brain wt × 100/ body wt), hepatosomatic index (HSI; liver wt × 100/body wt), and gonadosomatic index (GSI; gonad wt × 100/body wt) were calculated. The zebrafish were maintained in accordance with guidelines for the care and use of laboratory animals of the National Institute for Food and Drug Control of China. All animals were treated humanely and with the aim of alleviating any suffering. On the last day of the fish exposure, the fertilized eggs were collected from each tank. Then 100 eggs were randomly selected per each tank and were separately cultured in glass dishes containing fresh water without PHE until 6 dpf. Hatchability, survival, and malformation rate were determined from this F1 generation. The developmental status of zebrafish was observed with a microscope (Nikon, Japan).

2. Materials and methods 2.1. Test chemicals and instrumental analysis PHE (CAS No: 85-01-8) was obtained from Sigma-Aldrich (St. Louis, MO, USA). PHE was dissolved in dimethyl sulfoxide (DMSO) and the final DMSO concentration in the exposure water was 0.005% (v/v). To measure actual concentrations of the exposure media, water samples were collected from each tank for 4 times during the chronic exposure period. The samples were analyzed for PHE as described previously (Ma et al., 2010). Briefly, the water samples were passed through 0.7-μm glass-fiber filters (Whatman) and were kept at 4 °C until analysis. PHE was extracted from water samples using preconditioned C18 solid-phase extraction (SPE) cartridge (6 mL, 500 mg, Supelco, USA) with 10 mL n-hexane, followed by 10 mL methanol and 10 mL ultrapure water. Then, water sample was loaded at the flow rate of 2 mL/min. After drying the SPE cartridge, the object retained on the cartridge was eluted by 15 mL dichloromethane at the flow rate of 1 mL/min. The eluate was evaporated to dryness under a gentle stream of nitrogen at room temperature, and then dissolved in 0.5 mL acetonitrile for HPLC analysis. PHE concentration was determined by an Agilent 1100 HPLC Analytic System, equipped with a G-1321A scaning fluorescence detector (Agilent Technologies). The injection volume was 10 μL and analytical chromatography was performed with a flow rate of 1 mL/min at 25 °C. Mixture of 70% acetonitrile and 30% water was used as a mobile phase. Excitation at 250 nm/emission at 364 nm were chosen for all subsequent studies. Quantitative determination was performed from the fluorescence data (peak area measurement) by using an external calibration curve method. Recoveries of PHE in water

2.3. Plasma vitellogenin (VTG) and sex hormones measurement After exposure, blood was collected from caudal vein of the fish using a glass capillary tube, and blood samples from 3 fish of the same sex were pooled for analysis (n = 4 replicates). The blood samples were centrifuged at 7000×g for 5 min at 4 °C and the supernatant was collected and stored at −80 °C. Plasma concentrations of VTG were quantified by use of an enzyme-linked immunosorbent assay (ELISA) kit (Cayman Chemical Company, Ann Arbor, MI, USA; Cat. No. 10004995), following the manufacturer's instructions. Purified VTG from zebrafish was used as a standard. The absorbance was read at 492 nm. The concentration of VTG in each sample was calculated with nonspecific binding correction and regression analysis, performed by log−log transformation of the data. The detection limits were 0.12 ng/mL. The intra- and interassay coefficients of variance (CV) were < 10% and 15%, respectively. For the determination of sex hormones, the plasma collected was extracted twice with diethyl ether. Briefly, the plasma 2

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samples (∼10 μL) were diluted to 400 μL with UltraPure water and extracted with 2 mL diethyl ether at 2000×g for 10 min at 4 °C. The solvent used to extract hormones was evaporated under a stream of nitrogen, and the residue was dissolved in 120 μL of ELISA buffer. 17βestradiol (E2; Cat No. 582251) and testosterone (T; Cat No. 582701) were measured with commercially available kits according to the manufacturer's instructions (Cayman Chemical Company, Ann Arbor, MI, USA). No significant cross-reactivity or interference was observed for any kit. The intra- and inter-assay CV for measurements of E2 and T were determined to be < 20% and < 15%, respectively.

Table 1 Nominal and measured concentrations of PHE in waters during 120 days of exposure. The water was sampled just after renewal (T0) and before renewal (T24). Data are the means of three replicate tanks (mean ± SEM). TEXPa

Sampling times 5 dpf

Norminal 0 0.2

2.4. Quantitative real-time PCR assay of mRNA

1.0

Samples of brain, gonads and liver of each fish were collected after exposure and preserved in RNAlater storage solution (QIAGEN) at −80 °C for subsequent mRNA expression of genes. Total RNA was isolated from the sample by use of RNeasy mini-kit (QIAGEN), and RNA samples (0.1 μg RNA of male and female brain, 0.1 μg RNA of male gonad, and 1 μg RNA of female gonad) were used for reverse transcription using iScript cDNA Synthesis Kit (BIORAD, Hercules, CA, USA). Quantitative real-time PCR was performed by use of the SYBR Green PCR kit (Toyobo) on ABI 7300 System (PerkinElmer Applied Biosystems, Foster City, CA, USA). The primer sequences of the selected genes were obtained using the online Primer 3 program (http://frodo. wi.mit.edu/) and are shown in Table S1. The PCR reaction comprised an initial denaturation step at 95 °C for 2 min, followed by 40 cycles at 95 °C for 15 s, 60 °C for 15 s, and 72 °C for 1 min. A melting temperature-determining dissociation step was performed at 95 °C for 15s, 60 °C for 1 min, and 72 °C for 15 s at the end of the amplification phase. The dissociation curve was used to check the specificity of PCR products. Prior to the transcriptional assay, we assessed the amplification efficiencies of primers and the transcriptional stability of five candidate genes (β-actin, 18s, elfa, gapdh, and rpl8) commonly used as reference genes in both male and female. Then β-actin was selected as the reference gene for transcriptional assay in male brains, female brains and ovaries, and the rpl8 gene was selected as the reference gene for female liver and male liver and testes. For quantification of PCR results, the threshold cycle (Ct) was determined for each reaction. Ct values for each gene of interest were normalized to the endogenous control gene by using the △△Ct method. Normalized values were used to calculate the degree of induction or inhibition expressed as a “fold change” compared to normalized control values.

5.0 a b

T0 T24 T0 T24 T0 T24 T0 T24

b

LOD LOD 0.17 0.09 0.88 0.66 4.68 4.29

± ± ± ± ± ±

0.02 0.01 0.09 0.06 0.27 0.19

30 dpf

60 dpf

90 dpf

LOD LOD 0.16 0.10 0.96 0.69 4.62 4.14

LOD LOD 0.16 0.09 0.91 0.68 4.57 4.14

LOD LOD 0.15 0.08 0.87 0.63 4.65 4.21

± ± ± ± ± ±

0.01 0.02 0.05 0.04 0.29 0.23

± ± ± ± ± ±

0.02 0.01 0.11 0.05 0.21 0.26

± ± ± ± ± ±

0.01 0.01 0.07 0.04 0.22 0.18

TEXP = Time of the experiment at each sampling period (0 h and 24 h). LOD: limit of detection.

3.2. Growth and somatic indices of the F0 generation Exposure of the F0 generation to PHE did not significantly affect hatching and malformation rates (Table S2). Likewise, no significant effect was observed on the survival and sex ratio in the F0 fish exposed to any concentration of PHE (Table S2), and the overall survival rates were over 80%. There was no significant effects on either body length, body weight or condition factor (K-factor) at the end of the exposure. However, exposure to higher concentrations of PHE (1.0 and 5.0 μg/L) resulted in a significant increase in the HSI in male fish (Table 2). The gonadsomatic index (GSI) was significantly inhibited in the males exposed to 5.0 μg/L of PHE, while there were no significant differences between the GSI values of the female fish that were exposed to PHE (Table 2). 3.3. Egg production and F1 generation effects In comparison with the egg production in the control group, the fecundity in the fish exposed to PHE showed a significant decreasing trend for all doses of the exposure range, and the decrease was observed as significant (20.2 ± 2.1 eggs/female per day) at the highest concentration of 5.0 g/L of (Fig. 1). The malformation, hatchability and survival rates of F1 generation of each group were also recorded. As shown in Table 3, there were no significant differences in the malformation rates of F1 fish. However, a significant decrease in hatchability rate was observed in the F1 larvae derived from parents exposed to 5.0 μg/L of PHE relative to the control. In addition, the survival rates of the 1.0 and 5.0 μg/L exposure group were significantly lower than the corresponding control after 6 dpf (Table 3).

2.5. Statistical analysis The normality of the data was analyzed using the KolmogorovSmirnov test, and if necessary, data were log-transformed to approximate normality. Homogeneity of variances was verified by Levene's test. All data were shown as means ± standard error (SEM). One-way analysis of variance (ANOVA) was applied to calculate statistical significance followed by Dunnett's test. All the analyses were conducted with SPSS statistical software version 13.0 (SPSS, Inc., Chicago, IL, USA). A p < 0.05 was considered statistically significant.

3.4. Plasma sex hormones and VTG levels Long-term exposure to PHE caused statistically significant effects on sex hormones in fish (Fig. 2). In females, the concentrations of plasma E2 was significantly decreased by 35.7% in 5.0 μg/L exposure groups compared to the corresponding control. In addition, the concentration of plasma T were significantly reduced by 20.3% and 40.1% in 1.0 and 5.0 μg/L exposure groups compared with the solvent controls (Fig. 2A). In males, exposure to 1.0 and 5.0 μg/L of PHE significantly reduced the plasma E2 levels by 31.5% and 38.6%, respectively (Fig. 2B). However, no significant differences were observed in the plasma T level of males exposed to PHE (Fig. 2B). The concentrations of VTG measured in plasma samples from female and male fish are shown in Fig. 3. In males, no significant difference was observed between the treatment groups and the control. However, in females, a concentration-dependent decrease of VTG was measured after PHE exposure, significantly so for 5.0 μg/L of PHE compared with

3. Results 3.1. Chemical analysis The actual concentrations of PHE were measured before (T24) and just after (T0) the renewal of the exposure solutions at different sampling time points. Table 1 show that the measured concentrations were close to the nominal values at T0 and decreased relative to the nominal concentrations at T24. Concentrations in the solvent-control were below the limit of detection. For simplification, all results are presented using the nominal concentrations. 3

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Table 2 Growth and somatic index in F0 fish after exposure to PHE for 120 daysa. PHE (μg/L )

Length (cm) Weight (g) K factor BSI (%) HSI (%) GSI (%)

Female

Male

0

0.2

1.0

5.0

0

3.68 ± 0.05 0.50 ± 0.03 1.01 ± 0.06 0.77 ± 0.08 2.88 ± 0.33 15.66 ± 1.28

3.78 ± 0.05 0.57 ± 0.03 1.02 ± 0.04 0.74 ± 0.10 3.21 ± 0.25 16.37 ± 1.43

3.72 ± 0.06 0.48 ± 0.03 0.91 ± 0.02 0.79 ± 0.07 3.02 ± 0.23 15.87 ± 1.24

3.78 ± 0.04 0.51 ± 0.03 0.94 ± 0.03 0.76 ± 0.05 3.15 ± 0.28 16.81 ± 1.08

3.73 0.42 0.79 1.00 1.32 1.07

0.2 ± ± ± ± ± ±

0.05 0.02 0.03 0.15 0.18 0.12

3.68 0.42 0.85 1.09 1.45 0.92

1.0 ± ± ± ± ± ±

0.04 0.03 0.03 0.08 0.19 0.08

3.73 0.41 0.79 1.16 1.72 0.95

5.0 ± 0.04 ± 0.02 ± 0.03 ± 0.04 ± 0.18* ± 0.09

3.82 0.43 0.77 1.12 1.76 0.81

± 0.04 ± 0.02 ± 0.02 ± 0.07 ± 0.28* ± 0.08**

* Indicates significant difference between treatment groups and the corresponding control group (p < 0.05). ** Indicates significant difference between treatment groups and the corresponding control group (p < 0.01). a Mean ± standard error of the mean for 9 replicate fish.

measured hepatic ERα and ERβ gene transcription, however, no significant difference was observed between PHE treatment groups and control groups.

4. Discussion In the present study, we demonstrate that exposure to low concentrations of PHE causes reproductive toxicity in zebrafish as indicated by decreased egg numbers, inhibited plasma levels of VTG in females and disturbed sex hormones as well as altered gene transcription involved in steroidogenesis in the F0 generation. These results support the findings of a recent study conducted in medaka (Sun et al., 2015). Moreover, decreased hatchability and survival rate in the F1 generation were also observed in this study. These results imply that steroidogenic pathway could be responsive for endocrine disruption of PHE and might contribute to reproduction impairment in fish. It has been well known that sex steroid hormones is one of the most integrative and functional endpoints for reproduction. In the present study, PHE exposure significantly changed the plasma levels of sex hormones in both sexes. Specifically, in female fish, the plasma levels of E2 and T were significantly reduced. This result is consistent with a previous study demonstrating reduced plasma E2 levels in flounder (Platichthys flesus) after fed with PHE-contaminated food for 12 weeks (Monteiro et al., 2000a). Likewise, decreased plasma E2 and T levels were also observed in female zebrafish exposed to BaP (Hoffmann and Oris, 2006). In male fish, the concentrations of E2 were significantly decreased while production of T was not significantly altered. This observation is consistent with a previous study on long-term exposure of Sebastiscus marmoratus to PHE (Sun et al., 2011). In addition, increasing studies conducted in vitro and in vivo have demonstrated decreased plasma E2 in the presence of PAHs (Monteiro et al., 2000a, 2000b; Nicolas, 1999). These alterations in hormone levels suggest that PHE could disrupt the balance of steroid hormones in fish and could result in adverse effects on reproduction. The vertebrate reproductive system is under the control of the hypothalamic-pituitary-gonadal (HPG) axis, which operates through a hormonal cascade where production of gonadotropin-releasing hormone (GnRH) in the hypothalamus. GnRH regulates the synthesis and secretion of pituitary gonadotropins including follicle-stimulating hormone (FSH) and luteinizing hormone (LH), which transported to the

Fig. 1. F0 female fish spawning was expressed in terms of egg numbers per female/day. The results are presented as mean ± SEM. *P < 0.05 indicates significant difference between the exposure groups and the solvent-control group.

the control (Fig. 3). 3.5. Gene transcriptional profiles along the HPG axis The mRNA expression of selected genes involved in the HPG axis and liver are shown in Table 4. In the females exposed to 5.0 μg/L PHE, the expression of GnRH2 and FSHβ in the brain was significantly upregulated by 2.21 and 1.58-fold, respectively (Table 4). In contrast, the transcription of FSHβ and LHβ in male fish exposed to 5.0 μg/L was significantly downregulated (1.59 and 1.32-fold inhibition, respectively). In the ovary, exposure to PHE significantly inhibited 17β-HSD gene transcription (1.39-fold) in the 5.0 μg/L exposure, while CYP19a gene transcription was significantly down-regulated by 1.49- and 1.79fold after 1.0 and 5.0 μg/L PHE exposure, respectively (Table 4). In the testes, a strong up-regulation of CYP11a1 gene transcription was observed in a concentration-dependent manner (1.53- and 1.86-fold in 1.0 and 5.0 μg/L exposure groups, respectively) after PHE treatment, however, the expression of CYP19a was significantly inhibited after exposure to 5.0 μg/L of PHE (Table 4). In the present study, we also

Table 3 Development parameters in the F1 generation derived from F0 exposed to PHE (0, 0.2, 1.0, and 5.0 μg/L) for 120 days. PHE (μg/L)

0

0.2

1.0

5.0

Hatchability rate (%) Malformation rate (%) Survival rate (%)

89.2 ± 1.85 3.0 ± 0.56 92.3 ± 0.95

85.4 ± 1.76 2.67 ± 0.33 89.8 ± 0.88

86.8 ± 2.05 2.33 ± 0.67 85.2 ± 1.05*

72.1 ± 1.45* 3.67 ± 0.67 82.7 ± 2.16**

* Indicates significant difference between treatment groups and the corresponding control group (P < 0.05). ** Indicates significant difference between treatment groups and the corresponding control group (P < 0.01). 4

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Fig. 2. Plasma testosterone (T) and estradiol (E2) levels in female (A) and male zebrafish (B) after exposure to 0, 0.2, 1.0, and 5.0 μg/L of PHE for 4 months. Values represent the mean ± SEM of three replicate tanks. Significance between control and exposure groups are indicated by *P < 0.05, **P < 0.01.

Previous in vitro study demonstrated that PHE can interfere with steroidogenesis genes (Monteiro et al., 2000b). In this study, genes involved in steroidogenesis were also examined. Steroidogenic genes such as 17β-HSD and CYP19 encode enzymes that plays pivotal roles in the production of androgens and estradiol, respectively (Trant et al., 2001), and CYP11A1 is the first and rate-limiting enzyme that converts cholesterol into pregnenolone (Hsu et al., 2009). In the present study, we observed significant inhibition of the transcription of 17β-HSD and CYP19a in the ovary, which might partly, lead to the decreased concentration of T and E2 in females. In male fish, the upregulated CYP11A1 indicated a potential increase in the synthesis of sex steroids, and the down-regulation of CYP19a gene expression in testis could explain the reduction in the production of E2 in the presence of PHE. It has been well-known that VTG is hepatically synthesized in fish in response to stimulation by estrogenic chemicals. When estrogens enter the liver, they bind to ERs, forming ER complexes and then interact with estrogen response elements to regulate their expression (Watanabe et al., 2009). In addition, VTG induction is dependent on E2 concentration. Therefore, induction of VTG in males has been widely used as a biomarker for estrogenic compounds and is associated with reproductive disruption (Cheek et al., 2001). In the present study, we observed decreased plasma levels of VTG in females, accompanied with decreased E2 levels. Varied results have also been reported in previous studies. For instance, decreased plasma levels of E2 accompanied with induced hepatic vtg transcription have been demonstrated in other fish species upon chemicals exposure (Harris et al., 2001; Wang et al., 2013). Previous studies have indicated that induct ERs, especially ERa, appears to be a major mediator of VTG induction (Marlatt et al., 2008). However, in the livers of both sexes, no statistically significant difference in the mRNA expression level of ERα and ERβ was observed after PHE exposure. Thus, it is not clear that PHE exerts its effects through

Fig. 3. Plasma vitellogenin (VTG) levels in female and male zebrafish after exposure to 0, 0.2, 1.0, and 5.0 μg/L of PHE for 4 months. Values represent the mean ± SEM of three replicate tanks. Significance between control and exposure groups are indicated by *P < 0.05.

gonads and induce steroidogenesis producing sex steroid hormones, such as T and E2 (Nagahama and Yamashita, 2008; Sofikitis et al., 2008). Therefore, modulation of GnRHs by exposure to chemicals could disrupt production of gonadotropin hormones and subsequently have impact on reproduction. In the present study, transcriptions of GnRH2 and FSHβ were significantly up-regulated in female brain after exposure to PHE, which might be resulted from negative feedback of E2 on FSH. However, in male fish, inhibition of FSHβ and LHβ in brain after exposure to PHE was observed, suggesting possible delay in spermatogenesis and maturation.

Table 4 Gene transcription along the HPG axis in adult zebrafish exposed to PHE (μg/L) for 120 days.a,b Tissue

Gene

Female

Male

0 Brain

c

Gonad

Liver

gnrh2 fshβ lhβ 17β-hsd Cyp11a1 Cyp19a ERα ERβ

1.03 1.02 1.01 1.03 1.00 1.02 1.01 1.01

0.2 ± ± ± ± ± ± ± ±

0.06 0.09 0.09 0.03 0.02 0.11 0.01 0.03

1.21 1.12 1.05 0.95 1.01 0.86 0.92 1.12

1.0 ± ± ± ± ± ± ± ±

0.07 0.02 0.08 0.04 0.04 0.09 0.04 0.02

1.37 1.19 1.11 0.86 1.10 0.67 0.98 1.03

5.0 ± 0.21 ± 0.02 ± 0.12 ± 0.12 ± 0.03 ± 0.06* ± 0.04 ± 0.09

2.21 1.58 0.95 0.72 0.96 0.56 0.68 1.13

0 ± 0.23* ± 0.07** ± 0.18 ± 0.06* ± 0.06 ± 0.05** ± 0.07* ± 0.10

1.01 1.00 1.04 1.02 1.01 1.00 1.02 1.00

0.2 ± ± ± ± ± ± ± ±

0.02 0.03 0.08 0.03 0.02 0.04 0.04 0.03

1.03 0.94 1.00 1.05 0.90 0.82 0.91 1.00

1.0 ± ± ± ± ± ± ± ±

0.08 0.05 0.04 0.05 0.08 0.06 0.04 0.05

1.06 0.83 0.96 1.07 1.53 0.74 0.96 0.96

5.0 ± 0.07 ± 0.10 ± 0.04 ± 0.04 ± 0.21* ± 0.07* ± 0.13 ± 0.11

Significant differences between treatment groups and the corresponding control group are indicated by *P < 0.05 and **P < 0.01. a All results are expressed as mean ± SEM of three replicate tanks (two fish per replicate). b Gene transcription levels were expressed as fold change relative to the solvent control (DMSO). c The gene transcription levels were measured in the brain (including the hypothalamus and pituitary). 5

1.02 0.63 0.76 1.03 1.86 0.58 1.03 0.98

± 0.08 ± 0.05** ± 0.04* ± 0.04 ± 0.08** ± 0.05** ± 0.12 ± 0.11

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interaction with ERs, or induce estrogenic effects through other mechanisms in fish and the reasons for these observations need to be clarified in future. GSI is widely used as a biomarker of gonad condition in aquatic organisms and considered to be a useful indicator upon estrogens exposure (Harries et al., 1997). Exposure to synthetic estrogens such as 17a-ethynylestradiol has been shown to reduce GSI in male fish ((Kidd et al., 2007)). In this study, the GSI of the ovary showed no significant difference between the control and the treated groups, however, we observed a reduced GSI in exposed males. This result is consistent with a previous study, which demonstrated that exposure of male Sebastiscus marmoratus to 0.06 μg/L of PHE for 50 days significantly reduced the GSI of the testis. These observations support our hypothesis that PHE acts as an environmental estrogen on zebrafish gonad development. It has been well-known that endogenous estrogen may play a physiological role in spermatogenesis in male fish, and decrease of plasma E2 may result in the inhibition of spermatogenesis, subsequently reduce the GSI (Zhang et al., 2009). Combined with the plasma levels of sex hormones, the results in our study suggested that PHE mediated the reduction of GSI in male zebrafish might be attribute to the decreased plasma levels of E2. Conversely, in the present study, we observed significant increases in HSI in both sexes. Increased HSI values were also observed in zebrafish and carp treated with other estrogenic chemicals, such as fluorotelomer alcohols, nonylphenol, and ethinylestradiol (Liu et al., 2009; Schwaiger et al., 2000). The developmental toxicity of PHE exposure was evaluated in F0 and F1 generation embryos and larvae. PHE exposure did not affect F0 fish, however, decreased hatching and survival rates were observed in the F1 larvae after PHE exposure. These results demonstrate that longterm exposure to low concentration of PHE in parental fish caused adverse effects in the F1 generation. Additionally, our data suggest that offspring in the developmental stages are more sensitive than their parents who were exposed to PHE. In fish, parental exposure to estrogenic chemicals, especially after long-term exposure, such as diethylstilbestrol (Zhong et al., 2005), 17α-ethynylestradiol (Zha et al., 2008), polybrominated diphenyl ethers (Yu et al., 2011a), BaP (Corrales et al., 2014), bisphenol AF (Shi et al., 2015) or natural mixture of persistent organic pollutants (Lyche et al., 2013), have frequently been associated with developmental toxicity in F1 generation. Taken together, these results highlight the importance of studies in view of chronic exposure to environmental pollutants at low concentrations, as well as multigenerational assays to evaluate their environmental risks. In summary, the present study demonstrates that long-term exposure to low concentrations of PHE adversely affects reproduction of adult fish and impacts the development of offspring. PHE exposure could disrupt the balance of sex hormones by altering the expression of several key genes involved in steroidogenic pathway. The lower hatching and survival rate in the offspring of exposure group suggest a possibility of maternal exposure of PHE could increase the prevalence of adverse health signs in the offspring. However, given the fact that PHE exists with other PAHs or toxicants (e.g. heavy metals) in natural environment, future research on environmental risk assessment may be necessary to evaluate the chronic effects on reproduction in fish exposed to PHE with other environmental pollutants.

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