Reproduction impairment and endocrine disruption in adult zebrafish (Danio rerio) after waterborne exposure to TBOEP

Aquatic Toxicology 182 (2017) 163–171

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Reproduction impairment and endocrine disruption in adult zebrafish (Danio rerio) after waterborne exposure to TBOEP Qinglong Xu a,1 , Ding Wu b,1 , Yao Dang a , Liqin Yu a , Chunsheng Liu a,c,d , Jianghua Wang a,∗ a

College of Fisheries, Huazhong Agricultural University, Wuhan 430070, China Department of Urology, The Central Hospital of Wuhan, Tongji Medical College, Huazhong University of Science and Technology, Wuhan 430014, China Collaborative Innovation Center for Efficient and Health Productioon of Fisherise in Huhan Province, Changde 415000, China d Hubei Provincial Engineering Laboratory for Pond Aquaculture, Wuhan 430070, China b c

a r t i c l e

i n f o

Article history: Received 6 September 2016 Received in revised form 20 November 2016 Accepted 22 November 2016 Available online 24 November 2016 Keywords: Tris (2-butoxyethyl) phosphate Organophosphate flame retardants Reproduction impairment Endocrine disruption Zebrafish

a b s t r a c t Tris (2-butoxyethyl) phosphate (TBOEP) is widely used as a substitute of polybrominated diphenyl ethers (PBDEs). It has been frequently measured at concentrations of micrograms per liter (␮g/L) in surface waters and waste water. However, limited information is available about the reproduction toxicology of TBOEP. In this study, adult zebrafish pairs were exposed to TBOEP at concentrations of 0, 5, 50, and 500 ␮g/L for 21 days. The effects on reproduction, hormone concentration, transcription of genes along the hypothalamic-pituitary-gonadal (HPG) axis, and gonadal development were investigated. After exposure to TBOEP, plasma concentrations of 17␤-estradiol were significantly increased in both sexes of fish, while increase of testosterone was observed only in male fish. Transcription of genes along the HPG axis was significantly influenced by exposure to TBOEP in both male and female fish. Moreover, TBOEP decreases the average number of eggs production, as well as hatching success and survival rates in offspring. Histological examination shows inhibition of oocyte maturation in females and retardation spermiation in males, respectively. The results demonstrate that TBOEP could disturb the sex hormone balance by altering regulatory circuits of the HPG axis, affect gonadal development, eventually leading to disruption of reproductive performance and the development of progeny. © 2016 Elsevier B.V. All rights reserved.

1. Introduction Due to phased-out of main commercial polybrominated diphenyl ethers (PBDEs) mixtures, such as PentaBDE, the production and use of alternative flame retardants such as organophosphate flame retardants (OPFRs) have increased (Reemtsma et al., 2008). Among OPFRs, tris (2-butoxyethyl) phosphate (TBOEP) has been increasingly used in a number of applications and products as a substitute for PBDEs (McGee et al., 2012). Furthermore, TBOEP is used as a plasticizer in various products such as textiles, floor polish, varnish, plastics, furniture, foams, and electronic equipment (Marklund et al., 2003). It is an additive OPFR and is not chemically bind into final products. Thus, TBOEP can be discharged into the surrounding environment (Rodriguez et al., 2006). Previous studies have been demonstrated that TBOEP is frequently found

∗ Corresponding author at: College of Fisheries, Huazhong Agricultural University, Wuhan 430070, China. E-mail address: [email protected] (J. Wang). 1 These authors contributed equally to this work. http://dx.doi.org/10.1016/j.aquatox.2016.11.019 0166-445X/© 2016 Elsevier B.V. All rights reserved.

in waste water, effluent, surface water, ground water, soil, drinking water and even human milk (Fries and Puttmann, 2003; Cequier et al., 2014; Marklund et al., 2005; Andresen, 2006; Bacaloni et al., 2007; Stapleton et al., 2009; Sundkvist et al., 2010). In Sweden, the concentrations of TBOEP in influents and effluents of municipal waste water treatment plants were 35 ␮g/L and 30 ␮g/L, respectively, which demonstrated its resistance to waste water treatment processes (Marklund et al., 2005). In China, concentration of TBOEP in sediment has been reported to range from 1.00 to 5.00 mg/kg dm (dry mass) and it was the most prominent chemical in Tai Lake (Ch:Taihu) (Cao et al., 2012). In addition, a recent study reported that concentrations of TBOEP ranged from 0.07 to 3.50 ng/g wet weights in Lake trout (McGoldrick et al., 2014) and from 86.0 to ˜ et al., 2015). 98.4 ng/g dry weights in mullet fish (Alvarez-Munoz Although neurotoxicity, developmental toxicity, reproductive toxicity, and systemic effects of OPFRs have been reported (Veen and Boer, 2012), limited information is currently available on reproduction impairment and the underlying mechanisms of TBOEP. A recent report demonstrated that TBOEP increased concentrations of 17␤-estradiol (E2) and testosterone (T) by altering transcriptions of major steroidogenic genes in H295R cells (Liu et al., 2012). In

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contrast to this, after TBOEP treatment, the expression of genes relating to T synthesis and the concentration of T declined strongly in TM3 Leydig cells (Jin et al., 2015). Another study also reported that TBOEP induced pregnane X receptor agonistic activity in cell based on transactivation assays (Kojima et al., 2013). In fish, reproduction is regulated mainly by the hypothalamicpituitary-gonadal (HPG) axis (Ma et al., 2012). Theoretically, disrupting any point in the HPG axis can adversely affect the function of the reproductive endocrine system. It is well-known that OPFR families have endocrine disruption effect in fish. Tris (1, 3-dichloro-2-propyl) phosphate (TDCPP) and triphenyl phosphate (TPP) exposure resulted in a decline in fecundity, which was associated with significant increases of plasma E2 level and several transcriptions of steroidogenic genes (Liu et al., 2013). After TDCPP exposure, remarkable increases of plasma E2 and T levels in females, reduced egg quality, and increased malformation rates in the F1 generation have been observed (Wang et al., 2015). A recent study about TBOEP reported that exposure to TBOEP caused developmental malformations and changed the transcription of genes involved in the endocrine axis in zebrafish larva (Ma et al., 2016). Nevertheless, whether exposure to TBOEP can affect fish reproduction remains unknown. The present study was conducted to evaluate the impact of TBOEP on reproduction function in zebrafish. The influences on the sex steroid hormone level, mRNA transcription of vital genes in HPG axis, reproductive performance and gonadal histology in zebrafish were examined after 21 days exposure. This study will contribute to a better understanding of TBOEP reproductive toxicology in adult zebrafish and the relevant endocrine disruption potentials following exposure to TBOEP.

2. Material and methods

2.3. TBOEP exposure protocols Based on environmentally relevant concentrations of TBOEP (surface water 127 ng/L and wastewater up to 35 ␮g/L, respectively) (Cao et al., 2012), a gradient nominal concentrations were chosen (5, 50, and 500 ␮g/L, which are equivalent to 0.013, 0.13, and 1.3 ␮M, respectively). During exposure experiment, 8 male and 8 female mating fish were randomly selected and placed in each of three replicate tanks for each concentration of TBOEP for 21 d. Both control and treated groups received 0.01% (v/v) DMSO, which has no significant effects on development and reproduction in zebrafish (Han et al., 2014). During semi-static exposure, solutions were replaced for every 24 h with fresh carbon-filtered water containing corresponding concentrations of TBOEP. We have changed our description of exposure method in the revised version. 2.4. Evaluation of TBOEP on reproduction To examine the effects of TBOEP on reproduction, fish were spawned in different groups throughout the last 14 days of exposure. Spawned eggs were collected 2 h after the light turned on every morning, because the embryos developed normally and reached the blastula stage at 2 h post fertilization (hpf). The number of eggs per spawning event was recorded daily. Fecundity was reported as cumulative eggs per female during the last 14 days of exposure. Up to 100 fertilized eggs were randomly collected from each tank and separately cultured in glass dishes containing fresh water without TBOEP on the final day of spawning, and were observed for hatching rate and survival success at 5 dpf. Another fifteen randomly selected eggs from each tank were used to determine the egg diameter. The egg diameter was evaluated on a Leica M205FA microscope with a digital camera and software system, and the digital image was analyzed using LAS V4.5 software.

2.1. Zebrafish maintenance 2.5. Evaluation of adult zebrafish exposure endpoint parameters Wild-type adult male and female zebrafish (18-week old, AB wild-type) were obtained from zebrafish breeding center in Institute of Hydrobiology, Chinese Academy of Sciences (Wuhan, China) and were acclimated for approximately 7 days in a temperature controlled room (28 ◦ C ± 1 ◦ C). Male and female fish were cultured separately in glass tanks filled with dichlorinated tap water (pH 7.0–7.4). The culture water was renewed every 24 h. Fish were maintained under a photoperiod of 14:10 h light: dark and fed with fairy shrimp (Tianjin Red Sun Aquaculture Co., Ltd., Tianjin, China) three times a day. Water quality parameters, such as pH, conductivity, temperature, and dissolved oxygen, were measured weekly. All procedures were approved by the Institutional Animal Care and Use Committee (IACUC) of Huazhong Agriculture University for laboratory animal use, Wuhan of China. Culturing and breeding of fish was performed according to the common OECD protocol for fishes.

2.2. Chemicals reagents TBOEP (CAS NO.: 78-51-3; purity: 94%, product of Japan) was purchased from Sigma-Aldrich. A stock solution of TBOEP was prepared in dimethyl sulfoxide (DMSO), and diluted with DMSO to final concentrations immediately before use. The final concentration of solvent (DMSO) in test solutions did not exceed 0.01%. Test medium was prepared in fish culture water which is filtered and aerated (>24 h) tap water. TRIzol reagent and reverse transcription and SYBR Green kits were from Takara (Dalian, Liaoning, China), hormone detection kits were from Cayman Chemical Company (Ann Arbor, MI). All the reagents used in this study were of analytical grade.

After the 21 d of exposure, all fish in the mating tank were anesthetized on ice, and body weights and snout-vent lengths were measured. In each group, eight fish (4 male and 4 female) were selected randomly from each tank and the gonads were removed and weighed to determine the GSI. Then the gonads were prepared for sex hormones determination. Following Vitale et al. (2006), gonadosomatic index (GSI = 100 × [gonad weight (g)/body weight (g)]), hepatosomatic index (HSI = 100 × [hepar weight (g)/body weight (g)]), brain index (BSI = 100 × [brain weight (g)/body weight(g)]), and condition factor (K = 100 × [body weight (g)/total length3 (cm)]) values were calculated. 2.6. Sex hormones measurements After exposure, blood samples were collected as described by Liu et al. (2009). Briefly, samples of 4–10 ␮L of blood were collected from the caudal vein of each fish, and blood samples from 4 fish of the same sex were pooled as one replicate. The blood samples were centrifuged at 3500g for 10 min at 4 ◦ C, and the supernatant (females 8 ␮L; males 6 ␮L) was collected and stored at −80 ◦ C until analysis. Before ELISA could be performed, free steroids were extracted from the samples by using a previously reported method (Yu et al., 2014). Briefly, each blood sample was diluted to 490 ␮L with Milli-Q water in clean glass tubes, 2 mL diethyl ether was added to each glass and vortexed. The supernatants were removed and transferred to a clean tube. Repeat the extraction process three times. The collected supernatants were evaporated to dryness by heating to 30 ◦ C under a gentle stream of nitrogen and then stored at −80 ◦ C. The dried extracts were resuspended in 0.5 mL ELISA buffer.

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The T and E2 contents in adult zebrafish were measured with ELISA kits (Cayman Chemical Company; detection limits 6 pg/mL for T and 15 pg/mL for E2).

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Table 1 Quantitative polymerase chain reaction (qPCR) primers sequences used for gene expression assays. Gene name

Accession No.

Description

Sequence (5 -3 )

2.7. Histological examination

ˇ-actin

NM 131031

Four fish (2 male and 2 female) were selected randomly from each tank for histology observation in every concentration of TBOEP. After 21 d of exposure to TBOEP, the gonadal tissues of female and male zebrafish were fixed in paraformaldehyde’s solution for 24 h and then transferred to 70% ethanol. After dehydration in ethanol, the tissues were embedded in paraffin wax, and the sections were cut at 5 ␮m thickness and stained with hematoxylin and eosin. Two cross-sections of the gonads per fish from each treatment were selected randomly and photographed using a Leica microscope (Leica M205FA). Quantitative staging of the ovaries and testes was performed as described by Shang et al. (2006) and Liu et al. (2010). Briefly, for male fish, 10 lobules were randomly selected for examination, and 10 sections from each lobule were examined for each fish. The number of sections showing distinct stages of development was recorded: for spermatocyte development, the stages were spermatogonia, spermatocytes, and spermatids (Johnson et al., 2009). For female fish, oocytes were examined and classified into 4 developmental stages: oocyte/oogonia, previtellogenic, vitellogenic, and preovulatory oocytes (Johnson et al., 2009; Villeneuve et al., 2010). The proportion of each cell type at each stage was expressed as a percentage of the total number of follicles in the section examined.

gnrh2

AY657018

gnrh3

NM 182887

gnrhr1

NM 001144980

gnrhr2

NM 001144979

gnrhr3

NM 001177450

gnrhr4

NM 001098193

fshˇ

NM 205624

lhˇ

NM 205622

cyp19b

AF183908

er˛

NM 152959

erˇ

NM 174862

ar

NM 001083123

fshr

NM 001001812

lhr

AY424302

cyp17

AY281362

cyp19a

AF226620

cyp11a

NM 152983

star

NM 131663

hmgra

BC155135

hmgrb

NM 001014292

3ˇhsd

AY279108

17ˇhsd

AY306005

vtg1

AF406784

vtg3

AF254638

Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse Forward Reverse

TGCTGTTTTCCCCTCCATTG TCCCA TGCCAACCATCACT CTGAGACCGCAGGGAAGAAA TCACGAATGAGGGCATCCA TTGCCAGCACTGGTCATACG TCCATTTCACCAACGCTTCTT ACCCGAATCCTCGTGGAAA TCCACCCTTGCCCTTACCA CAACCTGGCCGTGCTTTACT GGACGTGGGAGCGTTTTCT GAGGCGCAGCGGAACA GTCATCTTCAGCGTCTTCATCCT CACCAACAACAAGCGCAAGT GGCAACGGTGAGGTTCATG GCTGTCGACTCACCAACATCTC GTGACGCAGCTCCCACATT GGCTGCTCAGAGCTTGGTTT TCCACCGATACCGTCTCATTTA GTCGTTACTTCCAGCCATTCG GCAATGTGCTTCCCAACACA CAGACTGCGCAAGTGTTATGAAG CGCCCTCCGCGATCTT TTCACCCCTGACCTCAAGCT TCCATGATGCCTTCAACACAA TCTGGGTTGGAGGTCCTACAA GGTCTGGAGCGAAGTACAGCAT CGTAATCCCGCTTTTGTTCCT CCATGCGCTTGGCGATA GGCCATCGCCGGAAA GGTTAATTTGCAGCGGCTAGTG TCTTTGACCCAGGACGCTTT CCGACGGGCAGCACAA GCTGACGGATGCTCAAGGA CCACGATGCACCGCAGTA GGCAGAGCACCGCAAAA CCATCGTCCAGGGATCTTATTG GGTCTGAGGAAGAATGCAATGAT CCAGGTCCGGAGAGCTTGT GAATCCACGGCCTCTTCGT GGGTTACGGTAGCCACAATGA TGGCCGGACCGCTTCTA GTTGTTGCCATAGGAACATGGA AGGCACGCAGGAGCACATCT CCAATCGTCTTTCAGCTGGTAA TGCATCTCGCATCAAATCCA GTCCAAGTTCCGCATAGTAGCA TCCATTGCTGAAAACGACAA TGCATTCAGCACACCTCTCA GGTGGTTCTTGGACTTGGTT CACAGGAGAGGATGGGATTT

2.8. Quantitative real-time polymerase chain reaction assay Six females or male fish from each treatment were randomly sampled. The gonad and brain were collected and preserved in Trizol reagent at −80 ◦ C. Briefly, total RNA was isolated using Trizol regent and digested with RNase-free DNaseI (Promega) following the manufacturer’s instructions. Concentrations of total RNA were estimated by spectrophotometric analysis at 260 nm. The purity of the RNA in each sample was verified by determining the A260/A280 ratio and by confirming the purity of 1.0 ␮g RNA using 1% agaroseformaldehyde gel electrophoresis with ethidium bromide staining. The purified RNA was used immediately for reverse transcription (RT) or stored at −80 ◦ C until analysis. Synthesis of first-strand ® cDNA was performed using a PrimeScript RT Reagent Kit (TaKaRa) following the manufacturer’s instructions. Quantitative real-time polymerase chain reaction (PCR) was conducted on an Step One Plus Real-time System (Perkin-Elmer Applied Biosystem) using the ® SYBR GreenTM PCR kit (Toyobo). The primer sequences of the selected genes were obtained using the online Primer 3 program (http://frodo.wi.mit.edu/) and are listed in Table 1. The thermal cycle was set at 95 ◦ C for 2 min; this was followed by 40 cycles at 95 ◦ C for 15 s, 60 ◦ C for 15 s, and 72 ◦ C for 1 min and a final cycle of 95 ◦ C for 15 s, 60 ◦ C for 1 min, and 95 ◦ C for 15 s. The ˇ–actin gene was used as internal control, because mRNA expression of this gene in both male and female tissues was not affected by TBOEP exposure in the present study (data not shown). The mRNA expression level of each target gene was normalized to the mRNA content of its reference gene, and changes in the mRNA expression of the relevant genes were analyzed by using the 2−Ct method (Livak and Schmittgen, 2001). (ar: androgen receptor; cyp11a: cytochrome P450 sidechain cleavage; 3ˇhsd: 3␤-hydroxysteroid dehydrogenase; cyp17: cytochrome P450; 17ˇhsd: 17␤-hydroxysteroid dehydrogenase; cyp19: cytochrome P450 19; gnrh: gonadotropin-releasing hormone; gnrhr: gonadotropin-releasing hormone receptor; hmgr: hydroxymethylglutaryl CoA reductase; lhˇ: luteinizing hormone

␤; lhr: luteinizing hormone receptor; fshˇ: follicle stimulating hormone ␤; fshr: follicle stimulating hormone receptor; star: steroidogenic acute regulatory protein; vtg: vitellogenin) 2.9. Statistical analysis All statistical analyses were performed using SPSS 19.0 (SPSS, Chicago, IL). The normality of the data and the homogeneity of variances were analyzed with the Kolmogorov-Smirnov test and Levene’s test, respectively. Levels of gene transcription in tissues were expressed as fold changes relative to the solvent control. When the data satisfied the assumption of homogeneity of variances, one-way analysis of variance (ANOVA) and Tukey’s multiple range test were used to determine significant differences between control and exposure groups. Correlation between transcriptional profile and the percentage of Vit & PreO in female zebrafish were examined by Spearman correlation analyses with Bonferroni correction. The value p < 0.05 was considered statistically significant. All values were expressed as the mean ± standard error (SEM).

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Fig. 1. Effects of TBOEP on reproductive parameters in zebrafish breeding pairs after 21 days exposure. (A) Cumulative fecundity in F0 generation exposed to various concentrations of TBOEP (each tank contains eight pairs of fish). (B) Egg diameters of F1 generation (15 eggs per tank). (C) F1 generation survival (100 embryos per tank). (D) F1 generation hatching (100 embryos per tank). The values are expressed as mean ± SEM. The asterisk indicates a value significantly different (*p < 0.05 and **p < 0.01) from that of the control group (treated with 0.01% DMSO).

3. Results 3.1. Reproductive endpoints Almost no mortality was observed in any of the experimental concentrations during the exposure period. Cumulative egg production was significantly reduced from the 10th day until 21th day by exposure to 50 ␮g/L and 500 ␮g/L TBOEP relative to the unexposed controls in female fish (Fig. 1A), and egg diameter was also significantly decreased (Fig. 1B). In addition, remarkably decreases in hatching success and survival rate were observed as exposure concentrations increased at 4 dpf (Fig. 1C and D). 3.2. Toxicological endpoints of somatic indices in adult zebrafish There was no statistically significant difference of K observed in exposed male or female fish relative to control. HSI was observed significantly decrease in the 50 ␮g/L and 500 ␮g/L treatment groups in female fish, while BSI was observed significant decrease in male fish after exposed to 500 ␮g/L TBOEP. GSI was significantly decreased in 500 ␮g/L TBOEP treatment group regardless of sex different. That was to say, TBOEP could affect GSI in both sex fish greatly (Fig. 2A and B). 3.3. Effects on plasma sex hormone levels In female fish, plasma E2 levels were increased significantly after exposure to 500 ␮g/L TBOEP (Fig. 3A), while plasma T concentrations generally were decreased but it didn’t have significant change (Fig. 3B). These changes in sex hormone concentrations led to a significant decrease of T/E2 ratio at 500 ␮g/L of TBOEP (Fig. 3C). While in male fish, plasma E2 and T concentrations were increased strongly after exposure to 50 and 500 ␮g/L TBOEP (Fig. 3A and B). And T/E2 ratio in male fish was increased remarkably by exposure to highest concentration of TBOEP (Fig. 3C). 3.4. Transcriptional changes of selected genes along the HPG axis Transcription levels of selected genes along the HPG axis were significantly influenced by exposure to TBOEP (Fig. 4).

Fig. 2. Effects on gonadal-somatic index (GSI), hepatic-somatic index (HSI), brainsomatic index (BSI) and K factor in (E) female and (F) male fish (n = 10). The values are expressed as mean ± SEM. The asterisk indicates a value significantly different (*p < 0.05 and **p < 0.01) from that of the control group (treated with 0.01% DMSO).

3.4.1. Brain As shown in Fig. 4, the transcription of gnrh2, gnrh3, gnrhr2, gnrhr3 and gnrhr4 was significantly down-regulated in the brain of males and females after exposed to 500 ␮g/L TBOEP (Fig. 4). The expression of gnrhr1 was significantly inhibited by 500 ␮g/L TBOEP in females, but it wasn’t influenced by any concentration of TBOEP in males. The strong down-regulation of the transcriptional levels of fshˇ and lhˇ was induced by 500 ␮g/L TBOEP, however significant up-regulation of er˛ was observed in the same concentration of both sex fish in treatment groups compared with the control (Fig. 4). The expression of cyp19b was significantly up-regulated in any treatment group of female fish, and it was affected only in highest concentrations of TBOEP in males (Fig. 4).

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Fig. 3. Effects on serum (A) 17␤-estradiol (E2) and (B) testosterone (T) levels and (C) the ratio of T/E2 in zebrafish after exposure to 0, 5, 50 and 500 ␮g/L of TBOEP for 21 days. The results are mean ± SEM of five replicate samples (four fish per replicate). Significance between control and exposure groups are indicated by *p < 0.05, **p < 0.01.

Fig. 4. Gene expression profiles of the selected endocrine pathways along HPG axis in zebrafish exposed to 0, 5, 50 and 500 ␮g/L of TBOEP for 21 days. The results are mean ± SEM and expressed as fold change relative to the corresponding control. There were six replicate samples in each treatment.

3.4.2. Gonads Exposure to TBOEP led to significant transcriptional changes in both testis and ovaries in adult zebrafish (Fig. 4). Among the steroidogenic genes, transcriptional levels of star, hmgrb, cyp11a, cyp17, cyp19a, 3ˇhsd, and 17ˇhsd were significantly up-regulated in the 500 ␮g/L treatment groups in ovaries of female fish (Fig. 4). While in male fish, exposure to 500 ␮g/L TBOEP caused strongly up-regulation of gene transcription for star, hmgrb, cyp17, 3ˇhsd, 17ˇhsd in testis. But the expression of cyp19a was significantly down-regulated in 50 and 500 ␮g/L TBOEP treatment male groups (Fig. 4). Gene expression of fshr and lhr was not significantly altered in either male or female fish after exposure to any concentrations of TBOEP.

3.4.3. Liver In the liver, the mRNA level of the hepatic vitellogenin1 (vtg1), vitellogenin3 (vtg3), and estrogenic receptor ˛ (er˛) genes was examined. In females group, vtg1, vtg3 and er˛ gene transcription was significantly up regulated after exposure to 500 ␮g/L TBOEP (Fig. 4). However, no significant effect of TBOEP on the transcription of vtg1, vtg3 and er˛ gene was observed in male livers. 3.5. Histological examination Histological examination showed that the percentages of oogonia (Oo) and previtellogenic oocytes (PreV) were significantly increased in the 50 and 500 ␮g/L exposure groups, respectively

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Fig. 5. Effects of TBOEP on gonadal development in zebrafish. (A) a-d: Respresentative photomicrographs of HE-stained ovary section. Oogonia (white arrow), previtellogenic (red star), vitellogenic (green triangle), and pre-ovulatory oocytes (black arrow) were indicated in female zebrafish. e-h: Respresentative photomicrographs of HE-stained testis section. Spermatogonia (white arrow), spermatocytes (red star), and spermatids (black arrow) were indicated in male zebrafish. (B) The percent ages of distinct stages in ovaries of female zebrafish. (C) The percentages of distinct stages in testis of male zebrafish. Values represent the mean ± SEM of four individual fish from three replicate tanks. Significance between control and exposure groups are indicated by *p < 0.05, **p < 0.01. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)

(Fig. 5A, a–d; Fig. 5B). The percentage of vitellogenic oocytes (Vit) and pre-ovulatory oocytes (PreO) were strongly decreased after 50 and 500 ␮g/L TBOEP exposure (Fig. 5B), respectively. In males, there was no significant difference in the percentage of spermatogonia (SPG) after TBOEP exposure (Fig. 5A, e–h; Fig. 5C), while a significant increase in the percentage of spermatocytes (SPC) was observed in 50 ␮g/L and 500 ␮g/L treated groups (Fig. 5C). However, the percentage of spermatid (SPD) was significantly decreased after 50 and 500 ␮g/L TBOEP exposure compared with the control (Fig. 5C). In addition, correlation between transcriptional profile and the percentage of Vit & PreO in female zebrafish were examined by Spearman correlation analyses with Bonferroni correction. The percentage of Vit was positively correlated with expressions of gnrhr4 in the brain (r > 0.5, p < 0.05) (Table 2). The percentage of PreO was positively correlated with expressions of gnrh3, gnrhr2, gnrhr3, erˇ, fshˇ and lhˇ in the brain(r > 0.5, p < 0.05). Meanwhile, the percentage of PreO was negatively correlated with expressions of star, 3ˇhsd and cyp17 in the ovary(r < 0.5, p < 0.05).However, no significant correlation was observed between the percentage of Vit & PreO and expressions of other tested genes in the HPG axis.

Table 2 Spearman Rank Correlation Coefficients and Probabilities between gene transcriptional profile and the percentage of Vit & PreO stage in ovary of female zebrafish. Spearman Rank Correlation Tissue

Gene

Vit

PreO

gnrh2 gnrh3 gnrhr1 gnrhr2 gnrhr3

0.056 0.229 0.350 0.174 0.304

0.487 0.536* 0.410 0.709** 0.643**

brain

gnrhr4 era erˇ ar fshˇ lhˇ cyp19b fshr lhr

0.633** 0.094 0.263 0.301 0.387 0.288 0.030 0.190 −0.246

0.370 −0.427 0.512* 0.472 0.746** 0.531* −0.300 −0.197 0.047

ovary

hmgra hmgrb star cyp11a 3ˇhsd cyp17 17ˇhsd cyp19a er˛

−0.236 0.230 −0.167 −0.466 −0.074 0.004 0.068 0.432 0.272

0.202 −0.340 −0.785** −0.410 −0.509* −0.682** −0.482 −0.441 0.001

liver

vtg1 vtg3 er˛

−0.006 −0.047 −0.052

−0.449 −0.325 −0.245

4. Discussion The present study demonstrated that waterborne exposure to TBOEP caused reproductive dysfunction in adult zebrafish, significantly disturbed plasma E2 and T levels, and altered transcriptional levels of steroidogenic genes along the HPG axis, which in turn affected gonadal development and adversely impaired the reproductive success in the parent. Although endocrine disrupting effects and reproduction impairment of TBOEP have been reported on other experimental subject previously (Jin et al., 2015), only limited

* **

Correlation is significant at the 0.05 level (two tailed). Correlation is significant at the 0.01 level (two tailed).

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Fig. 6. Schematic diagram and comparison of the expression alterations of genes in the brains, gonads and livers of female and male zebrafish induced by TBOEP. Red and green arrows indicate expression alterations of genes in female and male zebrafish, respectively. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)

information is available on zebrafish, that’s our original intention to conduct the experiment. A recent study reported that the PNEC of TBOEP in water was 24 ␮g/L (Ma et al., 2016), concentrations of TBOEP observed in aquatic systems are comparable or greater than values of PNEC. For example, the maximum concentration of TBOEP in effluent from wastewater treatment plants in Sweden exceeded the PNEC by 35 ␮g/L (Marklund et al., 2005). So it is likely that TBOEP at the current level of occurrence may cause direct damage to aquatic ecosystem, potential consequences of endocrine disruption following TBOEP exposure in aquatic environment is required for confirmation. Our results showed that exposure to TBOEP could cause reproductive dysfunction in adult zebrafish, which were manifested with significantly reduced cumulative egg production number, decreased egg diameter, hatching and survival rate, as well as immature gonadal development in zebrafish. All nutritional requirements for embryonic development are supplied by fish eggs. Thus, the decreased egg size may have influences on the nutritional content of the eggs, thereby leading to the decreased hatching, survival rate and delayed embryonic development. Our observation corresponds well with a similar report which evaluated developmental toxicity of TBOEP (Han et al., 2014). The present study revealed a significant decrease in the pre-ovulatory oocytes and vitellogenesis stage in ovary caused by 50 ␮g/L and 500 ␮g/L TBOEP. The results indicated a reduction of oocyte maturation, which thus may partly contribute to the decrease of GSI value. In female fish, the growth stage (vitellogenesis) is mainly under the control of FSH, while LH is involved in the final stages of oocyte maturation (Clelland and Peng, 2009). Here, the strong down-regulation of fshˇ and lhˇ gene in the brain was observed, suggesting inhibition of ovarian gametogenesis and oocyte maturation. The results of transcriptional level were agreed with our histological examination of a significantly decreased percentage of Vit and PreO stage in ovary. Significant decrease of the percentage of mature oocytes led to a decline in the number of cumulative eggs production after exposure

to TBOEP. These results suggested that TBOEP exposure could affect ovarian gametogenesis and induce reproductive toxicity. Similarly, FSH and LH are the most vital hormones controlling fish spermatogenesis in male fish (Schulz et al., 2010). The down-regulation of mRNA expression of lhˇ may result in a delay in the final stage of spermic maturation. Our histological examination demonstrated a significant reduction of mature SPD stage in testis. Exposure to TBOEP significantly altered plasma sex hormone levels as well as the transcription of genes involved in the steroidogenesis pathway. In the present study, plasma E2 levels were significantly increased in both sex fish, while plasma T levels were increased only in male fish. The measurement of sex hormones has been regarded as one of the most important biomarkers for reproduction inhibition in zebrafish (Ma et al., 2012; Liu et al., 2013; Ji et al., 2013). The change in sex steroid hormone concentrations may result in subsequent reproductive dysfunction. Aromatase (CYP19) enzymes are known to catalyze the final conversion of androgen to estrogen (Trant et al., 2001). In female fish, increased transcription of cyp19a in ovary may enhance the conversion of T to E2 in ovary. In addition, significant down-regulation of cyp19a gene in testis could explain the increase production of T in the presence of TBOEP in male fish. The ratio of T/E2 has been served as a sensitive biomarker of freak sex hormones in fish (Folmar et al., 1996; Orlando et al., 2004), and disequilibrating the balance between T and E2 could influence reproduction, sex development, and sex differentiation (Shang et al., 2006; Folmar et al., 1996). Exposure to TBOEP changed the ratio of T/E2 in both sexes of fish, which may impair fish sex differentiation, gonadal development and sex determination. In this study, liver vtg1 and vtg3 expression in female fish was significantly increased after exposed to 500 ␮g/L TBOEP, but no significant differences were found in male fish between the control and treatments. Vtg gene transcription is dependent on E2 concentration (Liu et al., 2010), and the increase of vtg gene was coincided with apparently increase trend of E2 in females following exposure to TBOEP.

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In fish, sex steroids hormones, including E2, manifest both positive and negative feedback at the level of the hypothalamus (long feedback loops) and pituitary (short feedback loops) (Blazquez et al., 1998; Trudeau, 1997; Veldhuis et al., 2009). Therefore, further studies need to explore control mechanisms about TBOEP-induced reproduction toxicology at the systems level (e.g., HPG axis). The alterations of the expression alteration of genes in the brains, gonads and livers of female and male zebrafish induced by TBOEP are summarized in a schematic diagram (Fig. 6). Steroidogenesisrelated genes expression after exposure to TBOEP corresponded well with variation in sex hormone levels. Steroidogenic genes such as star, hmgrb, cyp11a, cyp17, 3ˇhsd and 17ˇhsd were up-regulated in testis in male fish, which corresponded well with the strong increase of T concentrations induced by TBOEP. Similar increased gene expressions and E2 levels were observed in female fish. On the other hand, the transcriptional level of the gnrhs, gnrhrs, fshˇ and lhˇ was decreased after exposure to TBOEP. Generally, sufficient E2 content could initiate a negative feedback mechanism that would inactive the responsiveness of the pituitary to inhibition, leading to an decreased secretion of GnRHs and gonadotropins (Shupnik, 1996; Zhang et al., 2008). Increased levels of plasma E2 and T may be account for the down-regulation of these genes through negative feedback mechanisms to inhibit the gene expression of gnrhs, fshˇ and lhˇ. The positive feedback regulation of TBOEP could be partly overcome by negative feedback indicates that the potential endocrine disrupting of risks of TBOEP should not be neglected (Hou et al., 2016). Our study also demonstrated that parental exposure to TBOEP, would affect the normal development of offspring in a similar pattern as has been reported for other OPFRs (Liu et al., 2013; Wang et al., 2015). However, the underlying mechanism is unknown due to the shortage of a detailed understanding of the transfer process. Parental transfer mechanisms maybe also the underlying reason for the reproductive toxicity exposure to TBOEP. Further investigations are needed to confirm the precise mechanisms of parental transfer after exposure to TBOEP. 5. Conclusions In summary, our results clearly showed that short-term (21 days) waterborne exposure to TBOEP impair reproductive capacity and disrupt reproductive success in adults, altered plasma sex hormones levels by influencing the transcription of several key steroidogenic genes along the HPG axis. Future work may be necessary to study the reproduction toxicology of TBOEP in a systematic manner after the long term exposure of environmental relevant concentration of TBOEP. Conflict of interest statement The authors declare that they have no conflicts of interest. Acknowledgments This work was supported by the National Natural Science Foundation of China (31370525), the Fundamental Research Funds for the Central Universities (2014PY027, 2662015PY030) and the Natural Science Foundation of Hubei Province of China (2014CFA031). References ˜ D., Rodríquez-Mozaz, S., Maulvault, A.L., Tediosi, A., Alvarez-Munoz, Fernandez-Tejedor, M., Van den Heuvel, F., Kotterman, M., Marques, A., Barcelo, D., 2015. Occurrence of pharmaceuticals and endocrine disrupting compounds in macroalgaes bivalves, and fish from coastal areas in Europe. Environ. Res. 143, 56–64.

Andresen, J.A., 2006. Emission, Fate and Behaviour of Phosphor organic Flame Retardants and Plasticisers in the Aquatic Environment. Universität Duisburg-Essen, Fakultät für Chemie. Bacaloni, A., Cavaliere, C., Foglia, P., Nazzari, M., Samperi, R., Lagana, A., 2007. Liquid chromatography/tandem mass spectrometry determination of organophosphorus flame retardants and plasticizers in drinking and surface waters. Rapid Commun. Mass Spectrom. 21, 1123–1130. Blazquez, M., Bosma, P.T., Chang, J.P., Docherty, K., Trudeau, V.L., 1998. Gamma aminobutyric acid up-regulates the expression of a novel secretogranin-II messenger ribonucleic acid in the goldfish pituitary. Endocrinology 139, 4870–4880. Cao, S., Zeng, X., Song, H., Li, H., Yu, Z., Sheng, G., Fu, J., 2012. Levels and distributions of organophosphate flame retardants and plasticizers in sediment from Taihu Lake, China. Environ. Toxicol. Chem. 31, 1478–1484. Cequier, E., Ionas, A.C., Covaci, A., Marce, R.M., Becher, G., Thomsen, C., 2014. Occurrence of a broad range of legacy and emerging flame retardants in indoor environments in Norway. Environ. Sci. Technol. 48, 6827–6835. Clelland, E., Peng, C., 2009. Endocrine/paracrine control of zebrafish ovarian development. Mol. Cell. Endocrinol. 312, 42–52. Folmar, L.C., Denslow, N.D., Rao, V., Chow, M., Crain, D.A., Enblom, J., Marcino, J., Jr, L.J.G., 1996. Vitellogenin induction and reduced serum testosterone concentrations in feral male carp (Cyprinus carpio) captured near a major metropolitan sewage treatment plant. Environ. Health Perspect. 104, 1096–1101. Fries, E., Puttmann, W., 2003. Monitoring of the three organophosphate esters TBP, TCEP and TBEP in river water and ground water (Oder, Germany). J. Environ. Monit. 5, 346–352. Han, Z.H., Wang, Q.W., Fu, J., Chen, H.S., Zhao, Y., Zhou, B.S., Gong, Z.Y., Wei, S., Li, J., Liu, H.L., Zhang, X.W., Liu, C.S., Yu, H.X., 2014. Multiple bio-analytical methods to reveal possible molecular mechanisms of developmental toxicity in zebrafish embryos/larvae exposed to tris (2-butoxyethyl) phosphate. Aquat. Toxicol. 150, 175–181. Hou, J., Li, L., Wu, N., Su, Y.J., Lin, W., Li, G.Y., Gu, Z.M., 2016. Reproduction impairment and endocrine disruption in female zebrafish after long-term exposure to MC-LR: a life cycle assessment. Environ. Pollut. 208, 477–485. Ji, K., Hong, S., Kho, Y., Choi, K., 2013. Effects of Bisphenol S exposure on endocrine functions and reproduction of zebrafish. Environ. Sci. Technol. 47, 8793–8800. Jin, Y.X., Chen, G.L., Fu, Z.W., 2015. Effects of TBEP on the induction of oxidative stress and endocrine disruption in TM3 Leydig cells. Environ. Toxicol. 31, 1276–1286. Johnson, R., Wolf, J., Braunbeck, T., 2009. OECD Guidance Document for the Diagnosis of Endocrine-Related Histopathology of Fish Gonads. Organisation for Economic Co-operation and Development, Avialable from: www.oecd.org/ dataoecd/33/27/42140701.pdf. Kojima, H., Takeuchi, S., Itoh, T., Iida, M., Kobayashi, S., Yoshida, T., 2013. In vitro endocrine disruption potential of organophosphate flame retardants via human nuclear receptors. Toxicology 314, 76–83. Liu, C.S., Yu, L.Q., Deng, J., Lam, P.K.S., Wu, R.S.S., Zhou, B.S., 2009. Waterborne exposure to fluorotelomer alcohol 6:2 FTOH alter plasma sex hormone and gene transcription in the hypothalamic-pituitary-gonadal (HPG) axis of zebrafish. Aquat. Toxicol. 93, 131–137. Liu, C.S., Deng, J., Yu, L.Q., Ramesh, M., Zhou, B.S., 2010. Endocrine disruption and reproductive impairment in zebrafish by exposure to 8:2 fluorotelomer alcohol. Aquat. Toxicol. 96, 70–76. Liu, X.S., Ji, K., Choi, K., 2012. Endocrine disruption potentials of organophosphate flame retardants and related mechanisms in H295R and MVLN cell lines and in zebrafish. Aquat. Toxicol 114–115, 173–181. Liu, X.S., Ji, K., Jo, A., Moon, H.B., Choi, K., 2013. Effects of TDCPP or TPP on gene transcriptions and hormones of HPG axis, and their consequences on reproduction in adult zebrafish (Danio rerio). Aquat. Toxicol. 134–135, 104–111. Livak, K.J., Schmittgen, T.D., 2001. Analysis of relative gene expression data using real-time quantitative PCR and the 2 (Delta C(T)) method. Methods 25, 402–408. Ma, Y.B., Han, J., Guo, Y.Y., Lam, P.K.S., Wu, R.S.S., Giesy, J.P., Zhang, X.W., Zhou, B.S., 2012. Disruption of endocrine function in in vitro H295R cell-based and in in vivo assay in zebrafish by 2,4-dichlorophenol. Aquat. Toxicol. 106–107, 173–181. Ma, Z.Y., Tang, S., Su, G.Y., Miao, Y.Q., Liu, H.L., Xie, Y.W., Giesy, J.P., Saunders, D.M.V., Hecker, M., Yu, H.X., 2016. Effects of tris (2-butoxyethyl) phosphate (TBOEP) on endocrine axes during development of early life stages of zebrafish (Danio rerio). Chemosphere 144, 1920–1927. Marklund, A., Andersson, B., Haglund, P., 2003. Screening of organophosphorus compounds and their distribution in various indoor environments. Chemosphere 53, 1137–1146. Marklund, A., Andersson, B., Haglund, P., 2005. Organophosphorus flame retardants and plasticizers in Swedish sewage treatment plants. Environ. Sci. Technol. 39, 7423–7429. McGee, S.P., Cooper, E.M., Stapleton, H.M., Volz, D.C., 2012. Early zebrafish embryo genesis is susceptible to developmental TDCPP exposure. Environ. Health Perspect. 120, 1585–1591. McGoldrick, D.J., Letcher, R.J., Barresi, E., Keir, M.J., Small, J., Clark, M.G., Sverko, E., Backus, S.M., 2014. Organophosphate flame retardants and organosiloxanes in predatory freshwater fish from locations across Canada. Environ. Pollut. 193, 254–261.

Q. Xu et al. / Aquatic Toxicology 182 (2017) 163–171 Orlando, E.F., Kolok, A.S., Binzcik, G.A., Gates, J.L., Horton, M.K., Lambright, C.S., Gray, J.L.E., Soto, A.M., Jr, L.J.G., 2004. Endocrine-disrupting effects of cattle feedlot effluent on an aquatic sentinel species, the fathead minnow. Environ. Health Perspect. 112, 353–358. Reemtsma, T., Quintana, J.B., Rodil, R., Garcia-Lopez, M., Rodriguez, I., 2008. Organophosphorus flame retardants and plasticizers in water and air I. Occurrence and fate. Trend Anal. Chem. 27, 727–737. Rodriguez, I., Calvo, F., Quintana, J.B., Rubi, E., Rodil, R., Cela, R., 2006. Suitability of solid-phase micro extraction for the determination of organophosphate flame retardants and plasticizers in water samples. J. Chromatogr. A 1108, 158–165. Schulz, R.W., Franca, L.R., Lareyre, J., LeGac, F., Chiarini-Garcia, H., Nobrega, R.H., Miura, T., 2010. Spermatogenesis in fish. Gen. Comp. Endocrinol. 165, 390–411. Shang, E.H.H., Yu, R.M.K., Wu, R.S.S., 2006. Hypoxia affects sex differentiation and development, leading to a male-dominated population in zebrafish (Danio rerio). Environ. Sci. Technol. 40, 3118–3122. Shupnik, M.A., 1996. Gonadotropin gene modulation by steroids and gonadotropin releasing hormone. Biol. Reprod. 54, 279–286. Stapleton, H.M., Klosterhaus, S., Eagle, S., Fuh, J., Meeker, J.D., Blum, A., Webster, T.F., 2009. Detection of organophosphate flame retardants in furniture foam and US house dust. Environ. Sci. Technol. 43, 7490–7495. Sundkvist, A.M., Olofsson, U., Haglund, P., 2010. Organophosphorus flame retardants and plasticizers in marine and fresh water biota and in human milk. J. Environ. Monit. 12, 943–951. Trant, J.M., Gavasso, S., Ackers, J., Chung, B.C., Place, A.R., 2001. Developmental expression of cytochrome P450 aromatase genes (CYP19a and CYP19b) in zebrafish fry (Danio rerio). J. Exp. Zool. 290, 475–483.

171

Trudeau, V.L., 1997. Neuroendocrine regulation of gonadotrophin II release and gonadal growth in the goldfish, Carassius auratus. Rev. Reprod. 2, 55–68. Veen, I.V.D., Boer, J.D., 2012. Phosphorus flame retardants: properties, production, environmental occurrence, toxicity and analysis. Chemosphere 88, 1119–1153. Veldhuis, J.D., Keenan, D.M., Liu, P.Y., Iranmanesh, A., Takahashi, P.Y., Nehra, A.X., 2009. The aging male hypothalamic-pituitary-gonadal axis: pulsatility and feedback. Mol. Cell. Endocrinol. 229, 14–22. Villeneuve, D.L., Garcia-Reyero, N., Martinovic, D., Cavallin, J.E., Mueller, N.D., Wehmas, L.C., Kahl, M.D., Linnum, A., Perkins, E.J., Ankley, G.T., 2010. Influence of ovarian stage on transcript profiles in fathead minnow (Pimephales promelas) ovary tissue. Aquat. Toxicol. 98, 354–366. Vitale, F., Svedang, H., Cardinale, M., 2006. Histological analysis invalidates macroscopically determined maturity ogives of the Kattegat cod (Gadus morhua) and suggests new proxies for estimating maturity status of individual fish. ICES J. Mar. Sci. 63, 485–492. Wang, Q.W., Lam, J.C.W., Han, J., Wang, X.F., Guo, Y.Y., Lam, P.K.S., Zhou, B.S., 2015. Developmental exposure to the organophosphorus flame retardant tris (1,3-dichloro-2-propyl) phosphate: estrogenic activity, endocrine disruption and reproductive effects on zebrafish. Aquat. Toxicol. 160, 163–171. Yu, L.Q., Liu, C.S., Chen, Q., Zhou, B.S., 2014. Endocrine disruption and reproduction impairment in zebrafish after long-term exposure to DE-71. Environ. Toxicol. Chem. 33, 1354–1362. Zhang, X., Hecker, M., Park, J., Tompsett, A.R., Newsted, J., Nakayama, K., Jones, P.D., Newsted, J.L., Au, D.W.T., Kong, R.Y.C., Wu, R.S.S., Giesy, J.P., 2008. Real-time PCR array to study effects of chemicals on the hypothalamic-pituitary-gonadal axis of the Japanese medaka. Aquat. Toxicol. 88, 173–182.