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Effects of tris(2-butoxyethyl) phosphate exposure on endocrine systems and reproduction of zebrafish (Danio rerio)
Environmental Pollution 214 (2016) 568e574
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Environmental Pollution journal homepage: www.elsevier.com/locate/envpol
Effects of tris(2-butoxyethyl) phosphate exposure on endocrine systems and reproduction of zebrafish (Danio rerio)* Bareum Kwon a, b, 1, Hyejin Shin c, 1, Hyo-Bang Moon d, Kyunghee Ji a, **, Ki-Tae Kim c, * a
Department of Environmental Health, Graduate School of Yongin University, Yongin 17092, Republic of Korea CRI Global Institute of Toxicology, Croen Research Inc., Suwon 16614, Republic of Korea c Department of Environmental Engineering, Seoul National University of Science and Technology, Seoul 01811, Republic of Korea d Department of Marine Sciences and Convergent Technology, Hanyang University, Ansan 15588, Republic of Korea b
a r t i c l e i n f o
a b s t r a c t
Article history: Received 6 January 2016 Received in revised form 14 April 2016 Accepted 15 April 2016
Tris(2-butoxyethyl) phosphate (TBEOP), a widely used organophosphate flame retardant, has frequently been detected both in the environment and the biota. However, limited information is available on the effects of TBEOP on the endocrine system and its underlying mechanisms. We exposed adult zebrafish pairs to TBEOP at concentrations of 0, 2.1, 11, and 118 mg/L for 21 d, and investigated the effects on gene transcription and hormone production related to the hypothalamic-pituitary-gonadal (HPG) axis, and on reproduction. The adverse effects on the F1 generation were further examined. In male fish, plasma concentrations of 17b-estradiol were significantly increased along with up-regulation of cyp19a. Exposure to TBEOP at 118 mg/L led to a significant decrease in average egg production. Exposure of the F0 generation to TBEOP delayed hatching and lowered hatching rates in the F1 generation. The results demonstrate that exposure to TBEOP at environmentally relevant concentration levels could affect the sex hormone balance by altering regulatory circuits of the HPG axis, eventually leading to disruption of reproductive performance and the development of offspring. © 2016 Elsevier Ltd. All rights reserved.
Keywords: TBEOP Organophosphate flame retardants Endocrine disruption HPG axis Zebrafish
1. Introduction Flame retardants have been used widely in a variety of consumer products, such as textile fabrics, polyurethane foams, plastic materials, and electronic devices (Alaee et al., 2003). Polybrominated diphenyl ethers (PBDEs), organobromine compounds that are used as flame retardants, have received great attention due to their persistence in the environment and possible adverse effects on humans and wildlife (Wiseman et al., 2011). Consequently, penta-BDE and octa-BDE were added to the list of persistent organic pollutants (POPs) under the Stockholm Convention on POPs (UNEP, 2011). Along with the phase-out of major commercial PBDE mixtures from the markets of the United States and Europe,
*
This paper has been recommended for acceptance by Harmon Sarah Michele. * Corresponding author. Department of Environmental Engineering, Seoul National University of Science and Technology University, Seoul 01811, Republic of Korea. ** Corresponding author. E-mail address: [email protected] (K.-T. Kim). 1 Kwon and Shin contributed equally to this work and are listed in alphabetical order. http://dx.doi.org/10.1016/j.envpol.2016.04.049 0269-7491/© 2016 Elsevier Ltd. All rights reserved.
organophosphate flame retardants (OPFRs) have been alternatively increased worldwide (Covaci et al., 2011). Tris(2-butoxyethyl) phosphate (TBEOP), an OPFR, has been frequently detected in aquatic environments in many countries, including the United States (Benotti et al., 2009), Italy (Bacaloni et al., 2008), Spain (Gorga et al., 2015; Rodil et al., 2012), and China (Wang et al., 2011; Yan et al., 2012). For example, in water samples from Iberian rivers in Spain, TBEOP concentrations ranged from 5.3 to 659 ng/L (Gorga et al., 2015), and its maximum concentration in effluents from Swedish sewage treatment plants was 35 000 ng/L (Marklund et al., 2005). TBEOP is one of the most common OPFRs detected even in tap water in China with concentrations ranging from 24.1 to 151 ng/L, and boiling led to increased TBEOP concentrations in drinking water (Li et al., 2014). Recently, several studies have detected TBEOP in fish. For example, concentrations of TBEOP ranged from 86.0 to ~ oz et al., 2015), 98.4 ng/g dry weight in mullet fish (Alvarez-Mu n and from 0.07 to 3.50 ng/g wet weight in Lake trout (McGoldrick et al., 2014). While neurotoxicity, reproductive toxicity, and systemic effects have been reported for OPFRs (van der Veen and de Boer, 2012), information on endocrine disrupting effects and the underlying
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mechanisms of TBEOP is limited. Exposure to TBEOP increased concentrations of 17b-estradiol (E2) and testosterone (T) by altering transcriptions of major steroidogenic genes in the H295R cell line (Liu et al., 2012). In TM3 Leydig cells, the expression of genes related to T synthesis and the concentration of T declined remarkably after TBEOP treatment (Jin et al., 2015). Kojima et al. (2013) reported also that TBEOP induced pregnane X receptor agonistic activity in cellbased transactivation assays. In fish, reproduction is regulated primarily by the hypothalamicpituitary-gonadal (HPG) axis (Ma et al., 2012). Disruption at any point in the HPG axis can adversely affect the function of the reproductive endocrine system. It has been previously shown that adverse effects of OPFR families in fish are associated with endocrine disruption. Tris(1,3-dichloro-2-propyl) phosphate (TDCPP) and triphenyl phosphate (TPP) induced a decrease in fecundity, which was linked to significant increases of plasma E2 and transcriptions of several steroidogenic genes (Liu et al., 2013). Exposure to TDCPP induced significant increases of plasma E2 and T levels in females, reduced egg quality, and increased malformation rates in the F1 generation (Wang et al., 2015). Exposure to TDCPP, tricresyl phosphate (TCP), and TPP in zebrafish for 14 d caused a significant increase in plasma E2 concentration in both sexes, with upregulation of cyp19a genes (Liu et al., 2012). A recent study reported that exposure to TBEOP induced developmental malformations and altered the expression of genes involved in the endocrine axes in zebrafish larvae (Ma et al., 2016). However, very limited information is available on the effects on the reproductive system and underlying mechanisms of TBEOP in adult zebrafish. The present study was conducted to evaluate the endocrine disrupting effects of TBEOP on the HPG axis in zebrafish. The effects on the sex steroid hormone levels, mRNA expression of important genes in HPG axis, and reproductive performance in zebrafish were investigated after 21 d exposure. This study will help better understand endocrine disruption potentials in zebrafish and the related mechanisms following exposure to TBEOP. 2. Material and methods 2.1. Test chemicals and instrument analysis Tris(2-butoxyethyl) phosphate (TBEOP, CAS No. 78-51-3, purity 99%) was obtained from AccuStandard (New Haven, CT, USA). Solvent-free stock solutions were prepared by dissolving TBEOP in dichlorinated tap water, and diluted to final concentrations immediately before use. To measure the actual concentration of TBEOP in the exposure medium, we collected water samples with three replicates from each treatment group before (0 h) and after (48 h) exposure, and the water samples were stored at 80 C until chemical analysis. For the identification and quantification of TBEOP, chemical analysis was performed using a gas chromatograph interfaced with a mass spectrometer (GC/MS, Agilent 7890A/5975C MSD; Agilent Technologies, Wilmington, DE), using the method from a previous study (Liu et al., 2013). A 20 mL water sample was spiked with deuterated (d15) TPP (TPP-d15, purity 98%; Cambridge Isotope Laboratories, Andover, MA, USA) as a surrogate standard. After preparing in methanol (ultra-trace residue analysis grade, J.T. Baker, Phillipsburg, NJ, USA), the solution was extracted with 5 mL of dichloromethane (DCM) for 30 min on a mechanical shaker. The mixture was divided into water and DCM phases after standing for 20 min. The DCM phase was then transferred into a 15 mL polypropylene tube (BD Falcon, Franklin Lakes, NJ, USA), and the solutions were extracted twice with 5 mL of DCM. To remove any residual water, we added approximately 2 g of anhydrous sodium sulfate (pesticide residue analysis grade, Kanto Chemicals, Tokyo,
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Japan) to the DCM extracts. Then, we concentrated the extracts to approximately 1 mL with a stream of nitrogen gas. The GC/MS was operated under positive electron impact mode and ions were monitored using the selected ion monitoring mode (m/z 57 and 85). A DB5-MS capillary column (30 m length, 0.25 mm inner diameter, 0.25 mm film thickness; J&W Scientific, Palo Alto, CA, USA) was used for the separation of TBEOP. The recoveries of spiked TPP-d15 were 97± 17% (average ± standard deviation). Solvent (hexane) injected between the calibration standards showed negligible contamination or carryover. Procedural blanks were processed from each set of nine water samples to check for laboratory contamination. Blanks did not contain quantifiable amounts of TBEOP. The limit of detection for TBEOP (0.17 ng/mL) was calculated as three times the signal to noise ratio. Details about quantification with a GC/MS are shown in Table S1. The TBEOP concentrations throughout the manuscript are the measured concentrations. 2.2. Maintenance of zebrafish and TBEOP exposure Adult zebrafish (Danio rerio; AB strain) were cultured in 30-L glass tanks containing 25 L dechlorinated tap water at 26 ± 2 C under a photoperiod 16:8 h light/dark in the Molecular and Environmental Toxicology Laboratory at Yongin University (Yongin, Korea). The fish were fed with Artemia nauplii and mosquito larvae (Green Fish, Seoul, Korea) twice a day ad libitum. After feeding, the remaining food and feces were removed in 30 min. Eighteen greater than six month old mating pairs (i.e., nine male and nine female) per each treatment group were exposed to the control, 2.1, 11, and 118 mg/L TBEOP for 21 d with three replicates. The exposure was conducted with three mating pairs per glass tank at 26 ± 2 C with a photoperiod 16:8 h light/dark. We renewed the test solution three times per week, and fish were fed with commercially available Artemia nauplii and mosquito larvae twice a day. We recorded daily the mortality and number of eggs spawned. Water parameters (i.e., pH, conductivity, temperature, and dissolved oxygen) in the solution were routinely monitored. After 21 d exposure, all surviving fish were sacrificed using 0.1% 2phenoxyethanol (Sigma-Aldrich, St. Louis, MO, USA). Snout-vent length and total weight were measured to calculate the condition factor (K). Brain somatic index (BSI) and gonadosomatic index (GSI) were also calculated. For measurement of sex steroid hormones and mRNA expression analysis, four male and four female fish were randomly chosen from each treatment. On days 7 and 14 exposure, fertilized eggs were randomly selected from each treatment group and placed in 96-well plates (Corning Life Sciences, CA, USA) to analyze hatching rate, time to hatch, and larval survival in the F1 generation. Forty-eight eggs loaded into a 96-well plate were exposed to TBEOP (the same concentration) and others were exposed to clean water (control) for 6 d without solution renewal. 2.3. Hormone measurement After 21 d exposure, the tail of each adult zebrafish was transected using a glass capillary tube treated with heparin to collect blood from the caudal vein (Ji et al., 2013). Briefly, 5 mL of plasma collected from the female and male fish was centrifuged at 2000 rpm (i.e., 879g force) for 10 min, and stored at 80 C until further analysis. A plasma sample with 400 mL ultrapure water was extracted with 2 mL diethyl ether at 2100g for 10 min, and this procedure was repeated twice. To extract hormones, the solvent was evaporated, and the residues dissolved in 120 mL ELISA buffer. We quantified E2 (Cat No. 582251) and T (Cat No. 582701) concentrations by using an ELISA kit (Cayman Chemical Company, Ann
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Arbor, MI, USA) according to the manufacturer's instructions. The intra-assay coefficient of variations, which were determined at multiple points on the standard curves, was less than 10%.
males exposed to TBEOP, but significant statistical differences were not observed (Fig. 2B). TBEOP did not cause any significant effects on the ratio of E2 to T in either of the sexes (Fig. 2C).
2.4. Quantitative real-time polymerase chain reaction (qRT-PCR) analysis
3.3. mRNA expressions of genes in the HPG axis
Two tissues (brain and gonad) collected from each fish were preserved in 250 mL RNAlater reagent (QIAGEN, Korea, Ltd, Seoul, Korea) at 20 C until analysis. We selected several putative 22transcription factors that play critical roles in the HPG axis. Total RNA was extracted from the sample using the RNeasy mini kit (QIAGEN). The complementary DNA (cDNA) was synthesized from one mg total RNA using the iScript™ cDNA Synthesis Kit (BIORAD, Hercules, CA, USA) according to the manufacturer's instructions. Quantitative real-time PCR (qRT-PCR) was performed on an Applied Biosystems 7500 Real-Time PCR Instrument System (Applied Biosystems, Foster City, CA, USA) using 10 mL of 2 SYBR Green™ PCR master mix (Applied Biosystems). Samples were denatured at 95 C for 10 min, followed by 40 cycles of denaturation for 15 s at 95 C, annealing together with extension for 1 min at 60 C. For quantification of qRT-PCR results, the threshold cycle (Ct) was determined for each reaction. The relative Ct values for each gene of interest were normalized to a housekeeping gene (i.e., b-actin) expression using the 2△△Ct method. Housekeeping genes were chosen as the most stable genes in both the brain and gonad of male and female fish, using geNorm analysis, and details have been described in a previous study (Ji et al., 2013). Normalized values were used to calculate the degree of up-regulation or down-regulation expressed as a fold change compared to the normalized control values. A gene list was putatively selected and their accession numbers and primer sequences are shown in Tables S2 and S3, respectively. 2.5. Data analysis Normality and homogeneity of variances of observations were analyzed by the Shapiro-Wilk test and Levene's test, respectively. Differences between the control and exposure groups were evaluated using the Kruskal-Wallis test or one-way analysis of variance (ANOVA) using SPSS® version 18.0K for Windows® (SPSS, Chicago, IL, USA). The criterion for statistical significance was set at p < 0.05. To evaluate the dose-response relationship, a linear regression analysis was conducted. 3. Results 3.1. Toxicological endpoints of TBEOP in the F0 and F1 generations No lethality was observed in zebrafish exposed to any concentration of TBEOP. Condition factor (K), BSI, and GSI of adult zebrafish are summarized in Fig. 1A. Only the K in male fish was significantly decreased at 118 mg/L TBEOP, but other somatic indices showed no statistically significant differences from those of the control. The numbers of eggs per day per tank decreased in a concentrationdependent manner (Fig. 1B). Continuous exposure to 118 mg/L TBEOP resulted in a significant decrease in hatchability and larval survival (Fig. 1B). Parental exposure to TBEOP did not affect time to hatch (Fig. 1B). 3.2. Concentrations of sex hormones Exposure to TBEOP significantly increased plasma E2 in males at 118 mg/L TBEOP, while no significant changes were observed in females (Fig. 2A). Concentrations of plasma T gradually increased in
Exposure to TBEOP affected the transcriptional levels of selected genes involved in steroidogenesis (Fig. 3). In male fish, exposure to TBEOP induced up-regulation of gnrh2, gnrhr2, gnrhr4, and cyp19b in the brain (Fig. 3A), and 17bhsd and cyp19a in the testis (Fig. 3B). However, the transcriptions of gnrh3 and lhb in the brain (Fig. 3A) and hmgra in the testis (Fig. 3B) were significantly down-regulated in male zebrafish. In females, the significant up-regulation of brain fshb, cyp19b, era, and ar mRNAs (Fig. 3C), and ovary fshr, star, and cyp19a mRNAs (Fig. 3D) were observed. The transcription of hmgra and cyp11a was down-regulated when female zebrafish were exposed to 2.1, 11, and 118 mg/L TBEOP (Fig. 3D). 4. Discussion The present study demonstrates that the exposure to TBEOP caused reproductive dysfunction, and changed hormone concentrations and gene transcriptions in the HPG axis of zebrafish. Similar endocrine disrupting effects and reproduction impairment by other OPFRs have been previously reported. Concentrations of plasma E2 significantly increased in adult zebrafish exposed to TDCPP and TPP for 14 d (Liu et al., 2012). Significant decreases in fecundity along with significant increases in plasma E2 concentrations in male fish were observed after exposure to TDCPP and TPP (Liu et al., 2013). Exposure to TDCPP also induced a significant reduction in fecundity, which was accompanied by an altered sex hormone level and gene expression in the HPG axis (Wang et al., 2015). Exposure to 1,2,5,6-tetrabromocyclooctane, one of the brominated flame retardants (BFRs), resulted in a significant decrease in egg production in Japanese medaka with changes in the expression of genes in the HPG axis (Saunders et al., 2015). A significant decrease in the number of eggs was also reported in zebrafish exposed to the PBDE metabolite, DE-71, for 120 d at 5 ng/L (Han et al., 2013). The observations from these previous studies and our study suggest that reproduction impairment and estrogenic effects are common consequences of exposure to flame retardants. The measurement of sex hormones has been used as one of the most important biomarkers for reproduction inhibition in zebrafish (Ji et al., 2013; Liu et al., 2013; Ma et al., 2012), since the change in sex steroid hormone concentrations may cause subsequent reproductive dysfunction. In the present study, a significantly greater production of E2 in male zebrafish was observed after exposure to 118 mg/L TBEOP. Significant increases of E2 have also been previously reported in male fish exposed to TDCPP and TPP, but the effective concentrations were greater than or similar to our study, e.g., an increase of E2 was observed in zebrafish exposed to 1000 mg/L TDCPP and 200 mg/L TPP (Liu et al., 2013). Treatment with 200 mg/L TDCPP, TPP, and TCP also resulted in significantly greater concentrations of E2 in the plasma of male and female zebrafish (Liu et al., 2012). These results suggest that the potential of TBEOP for endocrine disruption is no less than that of other OPFRs. Interestingly, the adverse effects of TBEOP on hormone levels were sex dependent, with males being more sensitive than females. The underlying mechanism of sex dependence requires further investigation. The present study indicates that the exposure to TBEOP is associated with changes of gene regulation in the HPG axis. The gonadotropin releasing hormone (GnRH) is a central hormone for regulating the synthesis and release of gonadotropin hormones (GtHs) and also acts as a neuromodulator and regulates
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Fig. 1. Effects on somatic index and reproductive performance after exposure to tris(2-butoxyethyl) phosphate (TBEOP). (A) Condition factor and somatic index of F0 fish (n ¼ 9 for males and n ¼ 9 for females per each treatment group), (B) average number of eggs/breeding tank/day in F0 generation, and (C) the percentage of survival, hatchability, and time to hatch (d) in F1 generation. The values are mean ± standard deviation. Asterisk (*) indicates significant difference from control (p < 0.05).
Fig. 2. Effects of tris(2-butoxyethyl) phosphate on (A) 17b-estradiol (E2) concentration, (B) testosterone (T) concentration, and (C) E2/T ratio in plasma. The results are shown as mean ± standard deviation of four biological replicate samples. Asterisk indicates significant difference from control (p < 0.05).
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Fig. 3. Transcriptional response of 20 genes of hypothalamic-pituitary-gonad axis in male and female zebrafish after 21 d exposure to tris(2-butoxyethyl) phosphate. Responses in (A) male brain, (B) male gonad, (C) female brain, and (D) female gonad are summarized. mRNA expression is expressed as fold change compared to that of the control. The results are shown as mean ± standard deviation of four biological replicate samples. Asterisk (*) indicates significant difference between exposure groups and the control group (p < 0.05).
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reproductive behaviors (Bosma et al., 2000). The transcription of gnrh2, the most highly conserved form of GnRH in zebrafish (Gopinath et al., 2004), significantly increased in male zebrafish exposed to TBEOP, indicating that this compound could affect GnRHs in the brain. The observed disruption of gnrh2, gnrhr2, and gnrhr4 genes can be further related to subsequent disruption in GtHs and sex steroid hormones. These results are consistent with previous reports showing up-regulation of the gnrh2 gene in both male and female zebrafish when exposed to TDCPP, TPP, and DE-71 (Han et al., 2011; Liu et al., 2013). The responsiveness of target gonad tissues to the follicle-stimulating hormone (FSH) and luteinizing hormone (LH) is largely dependent on the expression of their specific receptors (i.e., FSHR and LHR). In the present study, transcriptions of fshb and fshr genes in females significantly increased after exposure to TBEOP at environmentally relevant concentrations, which could accelerate gametogenesis. These results are comparable with previous studies; the significant up-regulation of fshr gene in the ovary was observed in zebrafish exposed to TDCPP and TPP (Liu et al., 2013; Wang et al., 2015). The greater abundances of transcripts of fshr in ovary exposed to TBEOP might be a response to greater fshb released from the pituitary. Responses in steroidogenesis-related gene transcription after exposure to TBEOP corresponded well with changes in sex hormones. Aromatase (cyp19) enzymes are known to catalyze the final conversion of androgen to estrogen, and their activities are well correlated with the mRNA level of cyp19 (Trant et al., 2001). In the present study, an increase of E2 concentrations in male fish was accompanied by up-regulation of the aromatase genes (cyp19a and cyp19b). Increased transcription of both cyp19a in the testis and cyp19b in the brain may increase the conversion of T to E2. Similar patterns have been observed in male zebrafish exposed to TDCPP and TPP (Liu et al., 2013; Wang et al., 2015). Steroidogenic genes such as 17bhsd encode enzymes that participate in the basal synthesis of T. Increased concentrations of T in male zebrafish might be explained by the up-regulation of 17bhsd. The up-regulation of 17bhsd in zebrafish exposed to TBEOP in this study was in good agreement with a recent study by Ma et al. (2016). Parental exposure to endocrine disrupting chemicals is of concern because there might be no excretion mechanism in the eggs. In the present study, exposure to environmentally relevant concentrations of TBEOP in the F0 generation delayed hatching and reduced hatching rates in the offspring generation. Our observation corresponds well with Han et al. (2014) who evaluated developmental toxicity of TBEOP, showing that TBEOP exposure significantly reduced hatching rate, which was observed at much higher concentrations, i.e., two orders of magnitude, than those we used. Other recent studies have also reported the adverse effects of TDCPP on hatchability in zebrafish embryos (Liu et al., 2013; Wang et al., 2013). Despite the absence of a detailed understanding of the transfer process, the present study apparently indicates that parental exposure to low concentrations of TBEOP, which are currently used as one type of PBDE replacement, would affect the development of the F1 generation in a similar manner as has been reported for BFRs. 5. Conclusion Our results clearly show that exposure to TBEOP could affect the reproduction of zebrafish and development of progeny generations. The present study is the first report on the linkage of the transcription genes in the HPG axis with hormonal and reproductive changes in adult fish through exposure to TBEOP. It should be noted that such changes could take place even at environmentally relevant levels. The precise mechanisms of parental transfer and histopathological examination in gonad tissue deserve further
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investigation. Acknowledgement This study was supported by National Research Foundation of Korea (NRF; Project no. 2013R1A1A2012182). We would like to thank S. Jang, B.K. Kim, and N.K. Jung for technical assistance in toxicity testing. Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.envpol.2016.04.049. References € din, A., Bergman, Å., 2003. An overview of commercially used Alaee, M., Arias, P., Sjo brominated flame retardants, their applications, their use patterns in different countries/regions and possible modes of release. Environ. Int. 29, 683e689. ~ oz, D., Rodríquez-Mozaz, S., Maulvault, A.L., Tediosi, A., Fern Alvarez-Mu n andez , D., 2015. Tejedor, M., Van den Heuvel, F., Kotterman, M., Marques, A., Barcelo Occurrence of pharmaceuticals and endocrine disrupting compounds in macroalgaes, bivalves, and fish from coastal areas in Europe. Environ. Res. 143, 56e64. Bacaloni, A., Cucci, F., Guarino, C., Nazzari, M., Samperi, R., Lagana, A., 2008. Occurrence of organophosphorus flame retardant and plasticizers in three volcanic lakes of Central Italy. Environ. Sci. Technol. 42e46, 1898e1903. Benotti, M.J., Trenholm, R.A., Vanderford, B.J., Holady, J.C., Stanford, B.D., Snyder, S.A., 2009. Pharmaceuticals and endocrine disrupting compounds in US drinking water. Environ. Sci. Technol. 43e3, 597e603. Bosma, P.T., Rebers, F.E.M., van Dijk, W., Willems, P., Goos, H.J.T., Schulz, R.W., 2000. Inhibitory and stimulatory interactions between endogenous gonadotropinreleasing hormones in the Africa catfish (Clarias gariepinus). Biol. Reprod. 62, 731e738. Covaci, A., Harrad, S., Abdallah, M.A.E., Ali, N., Law, R.J., Herzke, D., de Wit, D.A., 2011. Novel brominated flame retardants: a review of their analysis, environmental fate and behavior. Environ. Int. 37, 532e556. Gopinath, A., Tseng, A.L., Whitlock, K.E., 2004. Temporal and spatial expression of gonadotropin releasing hormone (GnRH) in the brain of developing zebrafish (Danio rerio). Gene Expr. Patterns 4e1, 65e70. , D., 2015. Occurrence and spatial distribuGorga, M., Insa, S., Petrovic, M., Barcelo tion of EDCs and related compounds in waters and sediments of Iberian rivers. Sci. Total. Environ. 503e504, 69e86. Han, X.B., Lei, E.N.Y., Lam, M.H.W., Wu, R.S.S., 2011. A whole life cycle assessment on effects of waterborne PBDEs on gene expression profile along the brainpituitary-gonad axis and in the liver of zebrafish. Mar. Pollut. Bull. 63, 160e165. Han, X.B., Yuen, K.W.Y., Wu, R.S.S., 2013. Polybrominated diphenyl ethers affect the reproduction and development, and alter sex ratio of zebrafish (Danio rerio). Environ. Pollut. 182, 120e126. Han, Z., Wang, Q., Fu, J., Chen, H., Zhao, Y., Zhou, B., Gong, Z., Wei, S., Li, J., Liu, H., Zhang, X., Liu, C., Yu, H., 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, 175e181. 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, 8793e8800. Jin, Y., Chen, G., Fu, Z., 2015. Effects of TBEP on the induction of oxidative stress and endocrine disruption in TM3 Leydig cells. Environ. Toxicol. http://dx.doi.org/ 10.1002/tox.22137. 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 314e1, 76e83. Li, J., Yu, N., Zhang, B., Jin, L., Li, M., Hu, M., Zhang, X., Wei, S., Yu, H., 2014. Occurrence of organophosphate flame retardants in drinking water from China. Water Res. 54, 53e61. Liu, X., 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. 114e115, 173e181. Liu, X., 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. 134e135, 104e111. Ma, Y., Han, J., Guo, Y., Lam, P.K.S., Wu, R.S.S., Giesy, J.P., Zhang, X., Zhou, B., 2012. Disruption of endocrine function in in vitro H295R cell-based and in in vivo assay in zebrafish by 2,4-dichlorophenol. Aquat. Toxicol. 106e107, 173e181. Ma, Z., Tang, S., Su, G., Miao, Y., Liu, H., Xie, Y., Giesy, J.P., Saunders, D.M., Hecker, M., Yu, H., 2016. Effects of tris (2-butoxyethyl) phosphate (TBOEP) on endocrine axes during development of early life stages of zebrafish (Danio rerio). Chemosphere 144, 1920e1927. Marklund, A., Andersson, B., Haglund, P., 2005. Organophosphorus flame retardants and plasticizers in Swedish sewage treatment plants. Environ. Sci. Technol. 39,
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Effects of tris(2-butoxyethyl) phosphate exposure on endocrine systems and reproduction of zebrafish (Danio rerio)* Bareum Kwon a, b, 1, Hyejin Shin c, 1, Hyo-Bang Moon d, Kyunghee Ji a, **, Ki-Tae Kim c, * a
Department of Environmental Health, Graduate School of Yongin University, Yongin 17092, Republic of Korea CRI Global Institute of Toxicology, Croen Research Inc., Suwon 16614, Republic of Korea c Department of Environmental Engineering, Seoul National University of Science and Technology, Seoul 01811, Republic of Korea d Department of Marine Sciences and Convergent Technology, Hanyang University, Ansan 15588, Republic of Korea b
a r t i c l e i n f o
a b s t r a c t
Article history: Received 6 January 2016 Received in revised form 14 April 2016 Accepted 15 April 2016
Tris(2-butoxyethyl) phosphate (TBEOP), a widely used organophosphate flame retardant, has frequently been detected both in the environment and the biota. However, limited information is available on the effects of TBEOP on the endocrine system and its underlying mechanisms. We exposed adult zebrafish pairs to TBEOP at concentrations of 0, 2.1, 11, and 118 mg/L for 21 d, and investigated the effects on gene transcription and hormone production related to the hypothalamic-pituitary-gonadal (HPG) axis, and on reproduction. The adverse effects on the F1 generation were further examined. In male fish, plasma concentrations of 17b-estradiol were significantly increased along with up-regulation of cyp19a. Exposure to TBEOP at 118 mg/L led to a significant decrease in average egg production. Exposure of the F0 generation to TBEOP delayed hatching and lowered hatching rates in the F1 generation. The results demonstrate that exposure to TBEOP at environmentally relevant concentration levels could affect the sex hormone balance by altering regulatory circuits of the HPG axis, eventually leading to disruption of reproductive performance and the development of offspring. © 2016 Elsevier Ltd. All rights reserved.
Keywords: TBEOP Organophosphate flame retardants Endocrine disruption HPG axis Zebrafish
1. Introduction Flame retardants have been used widely in a variety of consumer products, such as textile fabrics, polyurethane foams, plastic materials, and electronic devices (Alaee et al., 2003). Polybrominated diphenyl ethers (PBDEs), organobromine compounds that are used as flame retardants, have received great attention due to their persistence in the environment and possible adverse effects on humans and wildlife (Wiseman et al., 2011). Consequently, penta-BDE and octa-BDE were added to the list of persistent organic pollutants (POPs) under the Stockholm Convention on POPs (UNEP, 2011). Along with the phase-out of major commercial PBDE mixtures from the markets of the United States and Europe,
*
This paper has been recommended for acceptance by Harmon Sarah Michele. * Corresponding author. Department of Environmental Engineering, Seoul National University of Science and Technology University, Seoul 01811, Republic of Korea. ** Corresponding author. E-mail address: [email protected] (K.-T. Kim). 1 Kwon and Shin contributed equally to this work and are listed in alphabetical order. http://dx.doi.org/10.1016/j.envpol.2016.04.049 0269-7491/© 2016 Elsevier Ltd. All rights reserved.
organophosphate flame retardants (OPFRs) have been alternatively increased worldwide (Covaci et al., 2011). Tris(2-butoxyethyl) phosphate (TBEOP), an OPFR, has been frequently detected in aquatic environments in many countries, including the United States (Benotti et al., 2009), Italy (Bacaloni et al., 2008), Spain (Gorga et al., 2015; Rodil et al., 2012), and China (Wang et al., 2011; Yan et al., 2012). For example, in water samples from Iberian rivers in Spain, TBEOP concentrations ranged from 5.3 to 659 ng/L (Gorga et al., 2015), and its maximum concentration in effluents from Swedish sewage treatment plants was 35 000 ng/L (Marklund et al., 2005). TBEOP is one of the most common OPFRs detected even in tap water in China with concentrations ranging from 24.1 to 151 ng/L, and boiling led to increased TBEOP concentrations in drinking water (Li et al., 2014). Recently, several studies have detected TBEOP in fish. For example, concentrations of TBEOP ranged from 86.0 to ~ oz et al., 2015), 98.4 ng/g dry weight in mullet fish (Alvarez-Mu n and from 0.07 to 3.50 ng/g wet weight in Lake trout (McGoldrick et al., 2014). While neurotoxicity, reproductive toxicity, and systemic effects have been reported for OPFRs (van der Veen and de Boer, 2012), information on endocrine disrupting effects and the underlying
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mechanisms of TBEOP is limited. Exposure to TBEOP increased concentrations of 17b-estradiol (E2) and testosterone (T) by altering transcriptions of major steroidogenic genes in the H295R cell line (Liu et al., 2012). In TM3 Leydig cells, the expression of genes related to T synthesis and the concentration of T declined remarkably after TBEOP treatment (Jin et al., 2015). Kojima et al. (2013) reported also that TBEOP induced pregnane X receptor agonistic activity in cellbased transactivation assays. In fish, reproduction is regulated primarily by the hypothalamicpituitary-gonadal (HPG) axis (Ma et al., 2012). Disruption at any point in the HPG axis can adversely affect the function of the reproductive endocrine system. It has been previously shown that adverse effects of OPFR families in fish are associated with endocrine disruption. Tris(1,3-dichloro-2-propyl) phosphate (TDCPP) and triphenyl phosphate (TPP) induced a decrease in fecundity, which was linked to significant increases of plasma E2 and transcriptions of several steroidogenic genes (Liu et al., 2013). Exposure to TDCPP induced significant increases of plasma E2 and T levels in females, reduced egg quality, and increased malformation rates in the F1 generation (Wang et al., 2015). Exposure to TDCPP, tricresyl phosphate (TCP), and TPP in zebrafish for 14 d caused a significant increase in plasma E2 concentration in both sexes, with upregulation of cyp19a genes (Liu et al., 2012). A recent study reported that exposure to TBEOP induced developmental malformations and altered the expression of genes involved in the endocrine axes in zebrafish larvae (Ma et al., 2016). However, very limited information is available on the effects on the reproductive system and underlying mechanisms of TBEOP in adult zebrafish. The present study was conducted to evaluate the endocrine disrupting effects of TBEOP on the HPG axis in zebrafish. The effects on the sex steroid hormone levels, mRNA expression of important genes in HPG axis, and reproductive performance in zebrafish were investigated after 21 d exposure. This study will help better understand endocrine disruption potentials in zebrafish and the related mechanisms following exposure to TBEOP. 2. Material and methods 2.1. Test chemicals and instrument analysis Tris(2-butoxyethyl) phosphate (TBEOP, CAS No. 78-51-3, purity 99%) was obtained from AccuStandard (New Haven, CT, USA). Solvent-free stock solutions were prepared by dissolving TBEOP in dichlorinated tap water, and diluted to final concentrations immediately before use. To measure the actual concentration of TBEOP in the exposure medium, we collected water samples with three replicates from each treatment group before (0 h) and after (48 h) exposure, and the water samples were stored at 80 C until chemical analysis. For the identification and quantification of TBEOP, chemical analysis was performed using a gas chromatograph interfaced with a mass spectrometer (GC/MS, Agilent 7890A/5975C MSD; Agilent Technologies, Wilmington, DE), using the method from a previous study (Liu et al., 2013). A 20 mL water sample was spiked with deuterated (d15) TPP (TPP-d15, purity 98%; Cambridge Isotope Laboratories, Andover, MA, USA) as a surrogate standard. After preparing in methanol (ultra-trace residue analysis grade, J.T. Baker, Phillipsburg, NJ, USA), the solution was extracted with 5 mL of dichloromethane (DCM) for 30 min on a mechanical shaker. The mixture was divided into water and DCM phases after standing for 20 min. The DCM phase was then transferred into a 15 mL polypropylene tube (BD Falcon, Franklin Lakes, NJ, USA), and the solutions were extracted twice with 5 mL of DCM. To remove any residual water, we added approximately 2 g of anhydrous sodium sulfate (pesticide residue analysis grade, Kanto Chemicals, Tokyo,
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Japan) to the DCM extracts. Then, we concentrated the extracts to approximately 1 mL with a stream of nitrogen gas. The GC/MS was operated under positive electron impact mode and ions were monitored using the selected ion monitoring mode (m/z 57 and 85). A DB5-MS capillary column (30 m length, 0.25 mm inner diameter, 0.25 mm film thickness; J&W Scientific, Palo Alto, CA, USA) was used for the separation of TBEOP. The recoveries of spiked TPP-d15 were 97± 17% (average ± standard deviation). Solvent (hexane) injected between the calibration standards showed negligible contamination or carryover. Procedural blanks were processed from each set of nine water samples to check for laboratory contamination. Blanks did not contain quantifiable amounts of TBEOP. The limit of detection for TBEOP (0.17 ng/mL) was calculated as three times the signal to noise ratio. Details about quantification with a GC/MS are shown in Table S1. The TBEOP concentrations throughout the manuscript are the measured concentrations. 2.2. Maintenance of zebrafish and TBEOP exposure Adult zebrafish (Danio rerio; AB strain) were cultured in 30-L glass tanks containing 25 L dechlorinated tap water at 26 ± 2 C under a photoperiod 16:8 h light/dark in the Molecular and Environmental Toxicology Laboratory at Yongin University (Yongin, Korea). The fish were fed with Artemia nauplii and mosquito larvae (Green Fish, Seoul, Korea) twice a day ad libitum. After feeding, the remaining food and feces were removed in 30 min. Eighteen greater than six month old mating pairs (i.e., nine male and nine female) per each treatment group were exposed to the control, 2.1, 11, and 118 mg/L TBEOP for 21 d with three replicates. The exposure was conducted with three mating pairs per glass tank at 26 ± 2 C with a photoperiod 16:8 h light/dark. We renewed the test solution three times per week, and fish were fed with commercially available Artemia nauplii and mosquito larvae twice a day. We recorded daily the mortality and number of eggs spawned. Water parameters (i.e., pH, conductivity, temperature, and dissolved oxygen) in the solution were routinely monitored. After 21 d exposure, all surviving fish were sacrificed using 0.1% 2phenoxyethanol (Sigma-Aldrich, St. Louis, MO, USA). Snout-vent length and total weight were measured to calculate the condition factor (K). Brain somatic index (BSI) and gonadosomatic index (GSI) were also calculated. For measurement of sex steroid hormones and mRNA expression analysis, four male and four female fish were randomly chosen from each treatment. On days 7 and 14 exposure, fertilized eggs were randomly selected from each treatment group and placed in 96-well plates (Corning Life Sciences, CA, USA) to analyze hatching rate, time to hatch, and larval survival in the F1 generation. Forty-eight eggs loaded into a 96-well plate were exposed to TBEOP (the same concentration) and others were exposed to clean water (control) for 6 d without solution renewal. 2.3. Hormone measurement After 21 d exposure, the tail of each adult zebrafish was transected using a glass capillary tube treated with heparin to collect blood from the caudal vein (Ji et al., 2013). Briefly, 5 mL of plasma collected from the female and male fish was centrifuged at 2000 rpm (i.e., 879g force) for 10 min, and stored at 80 C until further analysis. A plasma sample with 400 mL ultrapure water was extracted with 2 mL diethyl ether at 2100g for 10 min, and this procedure was repeated twice. To extract hormones, the solvent was evaporated, and the residues dissolved in 120 mL ELISA buffer. We quantified E2 (Cat No. 582251) and T (Cat No. 582701) concentrations by using an ELISA kit (Cayman Chemical Company, Ann
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Arbor, MI, USA) according to the manufacturer's instructions. The intra-assay coefficient of variations, which were determined at multiple points on the standard curves, was less than 10%.
males exposed to TBEOP, but significant statistical differences were not observed (Fig. 2B). TBEOP did not cause any significant effects on the ratio of E2 to T in either of the sexes (Fig. 2C).
2.4. Quantitative real-time polymerase chain reaction (qRT-PCR) analysis
3.3. mRNA expressions of genes in the HPG axis
Two tissues (brain and gonad) collected from each fish were preserved in 250 mL RNAlater reagent (QIAGEN, Korea, Ltd, Seoul, Korea) at 20 C until analysis. We selected several putative 22transcription factors that play critical roles in the HPG axis. Total RNA was extracted from the sample using the RNeasy mini kit (QIAGEN). The complementary DNA (cDNA) was synthesized from one mg total RNA using the iScript™ cDNA Synthesis Kit (BIORAD, Hercules, CA, USA) according to the manufacturer's instructions. Quantitative real-time PCR (qRT-PCR) was performed on an Applied Biosystems 7500 Real-Time PCR Instrument System (Applied Biosystems, Foster City, CA, USA) using 10 mL of 2 SYBR Green™ PCR master mix (Applied Biosystems). Samples were denatured at 95 C for 10 min, followed by 40 cycles of denaturation for 15 s at 95 C, annealing together with extension for 1 min at 60 C. For quantification of qRT-PCR results, the threshold cycle (Ct) was determined for each reaction. The relative Ct values for each gene of interest were normalized to a housekeeping gene (i.e., b-actin) expression using the 2△△Ct method. Housekeeping genes were chosen as the most stable genes in both the brain and gonad of male and female fish, using geNorm analysis, and details have been described in a previous study (Ji et al., 2013). Normalized values were used to calculate the degree of up-regulation or down-regulation expressed as a fold change compared to the normalized control values. A gene list was putatively selected and their accession numbers and primer sequences are shown in Tables S2 and S3, respectively. 2.5. Data analysis Normality and homogeneity of variances of observations were analyzed by the Shapiro-Wilk test and Levene's test, respectively. Differences between the control and exposure groups were evaluated using the Kruskal-Wallis test or one-way analysis of variance (ANOVA) using SPSS® version 18.0K for Windows® (SPSS, Chicago, IL, USA). The criterion for statistical significance was set at p < 0.05. To evaluate the dose-response relationship, a linear regression analysis was conducted. 3. Results 3.1. Toxicological endpoints of TBEOP in the F0 and F1 generations No lethality was observed in zebrafish exposed to any concentration of TBEOP. Condition factor (K), BSI, and GSI of adult zebrafish are summarized in Fig. 1A. Only the K in male fish was significantly decreased at 118 mg/L TBEOP, but other somatic indices showed no statistically significant differences from those of the control. The numbers of eggs per day per tank decreased in a concentrationdependent manner (Fig. 1B). Continuous exposure to 118 mg/L TBEOP resulted in a significant decrease in hatchability and larval survival (Fig. 1B). Parental exposure to TBEOP did not affect time to hatch (Fig. 1B). 3.2. Concentrations of sex hormones Exposure to TBEOP significantly increased plasma E2 in males at 118 mg/L TBEOP, while no significant changes were observed in females (Fig. 2A). Concentrations of plasma T gradually increased in
Exposure to TBEOP affected the transcriptional levels of selected genes involved in steroidogenesis (Fig. 3). In male fish, exposure to TBEOP induced up-regulation of gnrh2, gnrhr2, gnrhr4, and cyp19b in the brain (Fig. 3A), and 17bhsd and cyp19a in the testis (Fig. 3B). However, the transcriptions of gnrh3 and lhb in the brain (Fig. 3A) and hmgra in the testis (Fig. 3B) were significantly down-regulated in male zebrafish. In females, the significant up-regulation of brain fshb, cyp19b, era, and ar mRNAs (Fig. 3C), and ovary fshr, star, and cyp19a mRNAs (Fig. 3D) were observed. The transcription of hmgra and cyp11a was down-regulated when female zebrafish were exposed to 2.1, 11, and 118 mg/L TBEOP (Fig. 3D). 4. Discussion The present study demonstrates that the exposure to TBEOP caused reproductive dysfunction, and changed hormone concentrations and gene transcriptions in the HPG axis of zebrafish. Similar endocrine disrupting effects and reproduction impairment by other OPFRs have been previously reported. Concentrations of plasma E2 significantly increased in adult zebrafish exposed to TDCPP and TPP for 14 d (Liu et al., 2012). Significant decreases in fecundity along with significant increases in plasma E2 concentrations in male fish were observed after exposure to TDCPP and TPP (Liu et al., 2013). Exposure to TDCPP also induced a significant reduction in fecundity, which was accompanied by an altered sex hormone level and gene expression in the HPG axis (Wang et al., 2015). Exposure to 1,2,5,6-tetrabromocyclooctane, one of the brominated flame retardants (BFRs), resulted in a significant decrease in egg production in Japanese medaka with changes in the expression of genes in the HPG axis (Saunders et al., 2015). A significant decrease in the number of eggs was also reported in zebrafish exposed to the PBDE metabolite, DE-71, for 120 d at 5 ng/L (Han et al., 2013). The observations from these previous studies and our study suggest that reproduction impairment and estrogenic effects are common consequences of exposure to flame retardants. The measurement of sex hormones has been used as one of the most important biomarkers for reproduction inhibition in zebrafish (Ji et al., 2013; Liu et al., 2013; Ma et al., 2012), since the change in sex steroid hormone concentrations may cause subsequent reproductive dysfunction. In the present study, a significantly greater production of E2 in male zebrafish was observed after exposure to 118 mg/L TBEOP. Significant increases of E2 have also been previously reported in male fish exposed to TDCPP and TPP, but the effective concentrations were greater than or similar to our study, e.g., an increase of E2 was observed in zebrafish exposed to 1000 mg/L TDCPP and 200 mg/L TPP (Liu et al., 2013). Treatment with 200 mg/L TDCPP, TPP, and TCP also resulted in significantly greater concentrations of E2 in the plasma of male and female zebrafish (Liu et al., 2012). These results suggest that the potential of TBEOP for endocrine disruption is no less than that of other OPFRs. Interestingly, the adverse effects of TBEOP on hormone levels were sex dependent, with males being more sensitive than females. The underlying mechanism of sex dependence requires further investigation. The present study indicates that the exposure to TBEOP is associated with changes of gene regulation in the HPG axis. The gonadotropin releasing hormone (GnRH) is a central hormone for regulating the synthesis and release of gonadotropin hormones (GtHs) and also acts as a neuromodulator and regulates
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Fig. 1. Effects on somatic index and reproductive performance after exposure to tris(2-butoxyethyl) phosphate (TBEOP). (A) Condition factor and somatic index of F0 fish (n ¼ 9 for males and n ¼ 9 for females per each treatment group), (B) average number of eggs/breeding tank/day in F0 generation, and (C) the percentage of survival, hatchability, and time to hatch (d) in F1 generation. The values are mean ± standard deviation. Asterisk (*) indicates significant difference from control (p < 0.05).
Fig. 2. Effects of tris(2-butoxyethyl) phosphate on (A) 17b-estradiol (E2) concentration, (B) testosterone (T) concentration, and (C) E2/T ratio in plasma. The results are shown as mean ± standard deviation of four biological replicate samples. Asterisk indicates significant difference from control (p < 0.05).
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Fig. 3. Transcriptional response of 20 genes of hypothalamic-pituitary-gonad axis in male and female zebrafish after 21 d exposure to tris(2-butoxyethyl) phosphate. Responses in (A) male brain, (B) male gonad, (C) female brain, and (D) female gonad are summarized. mRNA expression is expressed as fold change compared to that of the control. The results are shown as mean ± standard deviation of four biological replicate samples. Asterisk (*) indicates significant difference between exposure groups and the control group (p < 0.05).
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reproductive behaviors (Bosma et al., 2000). The transcription of gnrh2, the most highly conserved form of GnRH in zebrafish (Gopinath et al., 2004), significantly increased in male zebrafish exposed to TBEOP, indicating that this compound could affect GnRHs in the brain. The observed disruption of gnrh2, gnrhr2, and gnrhr4 genes can be further related to subsequent disruption in GtHs and sex steroid hormones. These results are consistent with previous reports showing up-regulation of the gnrh2 gene in both male and female zebrafish when exposed to TDCPP, TPP, and DE-71 (Han et al., 2011; Liu et al., 2013). The responsiveness of target gonad tissues to the follicle-stimulating hormone (FSH) and luteinizing hormone (LH) is largely dependent on the expression of their specific receptors (i.e., FSHR and LHR). In the present study, transcriptions of fshb and fshr genes in females significantly increased after exposure to TBEOP at environmentally relevant concentrations, which could accelerate gametogenesis. These results are comparable with previous studies; the significant up-regulation of fshr gene in the ovary was observed in zebrafish exposed to TDCPP and TPP (Liu et al., 2013; Wang et al., 2015). The greater abundances of transcripts of fshr in ovary exposed to TBEOP might be a response to greater fshb released from the pituitary. Responses in steroidogenesis-related gene transcription after exposure to TBEOP corresponded well with changes in sex hormones. Aromatase (cyp19) enzymes are known to catalyze the final conversion of androgen to estrogen, and their activities are well correlated with the mRNA level of cyp19 (Trant et al., 2001). In the present study, an increase of E2 concentrations in male fish was accompanied by up-regulation of the aromatase genes (cyp19a and cyp19b). Increased transcription of both cyp19a in the testis and cyp19b in the brain may increase the conversion of T to E2. Similar patterns have been observed in male zebrafish exposed to TDCPP and TPP (Liu et al., 2013; Wang et al., 2015). Steroidogenic genes such as 17bhsd encode enzymes that participate in the basal synthesis of T. Increased concentrations of T in male zebrafish might be explained by the up-regulation of 17bhsd. The up-regulation of 17bhsd in zebrafish exposed to TBEOP in this study was in good agreement with a recent study by Ma et al. (2016). Parental exposure to endocrine disrupting chemicals is of concern because there might be no excretion mechanism in the eggs. In the present study, exposure to environmentally relevant concentrations of TBEOP in the F0 generation delayed hatching and reduced hatching rates in the offspring generation. Our observation corresponds well with Han et al. (2014) who evaluated developmental toxicity of TBEOP, showing that TBEOP exposure significantly reduced hatching rate, which was observed at much higher concentrations, i.e., two orders of magnitude, than those we used. Other recent studies have also reported the adverse effects of TDCPP on hatchability in zebrafish embryos (Liu et al., 2013; Wang et al., 2013). Despite the absence of a detailed understanding of the transfer process, the present study apparently indicates that parental exposure to low concentrations of TBEOP, which are currently used as one type of PBDE replacement, would affect the development of the F1 generation in a similar manner as has been reported for BFRs. 5. Conclusion Our results clearly show that exposure to TBEOP could affect the reproduction of zebrafish and development of progeny generations. The present study is the first report on the linkage of the transcription genes in the HPG axis with hormonal and reproductive changes in adult fish through exposure to TBEOP. It should be noted that such changes could take place even at environmentally relevant levels. The precise mechanisms of parental transfer and histopathological examination in gonad tissue deserve further
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investigation. Acknowledgement This study was supported by National Research Foundation of Korea (NRF; Project no. 2013R1A1A2012182). We would like to thank S. Jang, B.K. Kim, and N.K. Jung for technical assistance in toxicity testing. Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.envpol.2016.04.049. References € din, A., Bergman, Å., 2003. An overview of commercially used Alaee, M., Arias, P., Sjo brominated flame retardants, their applications, their use patterns in different countries/regions and possible modes of release. Environ. Int. 29, 683e689. ~ oz, D., Rodríquez-Mozaz, S., Maulvault, A.L., Tediosi, A., Fern Alvarez-Mu n andez , D., 2015. Tejedor, M., Van den Heuvel, F., Kotterman, M., Marques, A., Barcelo Occurrence of pharmaceuticals and endocrine disrupting compounds in macroalgaes, bivalves, and fish from coastal areas in Europe. Environ. Res. 143, 56e64. Bacaloni, A., Cucci, F., Guarino, C., Nazzari, M., Samperi, R., Lagana, A., 2008. Occurrence of organophosphorus flame retardant and plasticizers in three volcanic lakes of Central Italy. Environ. Sci. Technol. 42e46, 1898e1903. Benotti, M.J., Trenholm, R.A., Vanderford, B.J., Holady, J.C., Stanford, B.D., Snyder, S.A., 2009. Pharmaceuticals and endocrine disrupting compounds in US drinking water. Environ. Sci. Technol. 43e3, 597e603. Bosma, P.T., Rebers, F.E.M., van Dijk, W., Willems, P., Goos, H.J.T., Schulz, R.W., 2000. Inhibitory and stimulatory interactions between endogenous gonadotropinreleasing hormones in the Africa catfish (Clarias gariepinus). Biol. Reprod. 62, 731e738. Covaci, A., Harrad, S., Abdallah, M.A.E., Ali, N., Law, R.J., Herzke, D., de Wit, D.A., 2011. Novel brominated flame retardants: a review of their analysis, environmental fate and behavior. Environ. Int. 37, 532e556. Gopinath, A., Tseng, A.L., Whitlock, K.E., 2004. Temporal and spatial expression of gonadotropin releasing hormone (GnRH) in the brain of developing zebrafish (Danio rerio). Gene Expr. Patterns 4e1, 65e70. , D., 2015. Occurrence and spatial distribuGorga, M., Insa, S., Petrovic, M., Barcelo tion of EDCs and related compounds in waters and sediments of Iberian rivers. Sci. Total. Environ. 503e504, 69e86. Han, X.B., Lei, E.N.Y., Lam, M.H.W., Wu, R.S.S., 2011. A whole life cycle assessment on effects of waterborne PBDEs on gene expression profile along the brainpituitary-gonad axis and in the liver of zebrafish. Mar. Pollut. Bull. 63, 160e165. Han, X.B., Yuen, K.W.Y., Wu, R.S.S., 2013. Polybrominated diphenyl ethers affect the reproduction and development, and alter sex ratio of zebrafish (Danio rerio). Environ. Pollut. 182, 120e126. Han, Z., Wang, Q., Fu, J., Chen, H., Zhao, Y., Zhou, B., Gong, Z., Wei, S., Li, J., Liu, H., Zhang, X., Liu, C., Yu, H., 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, 175e181. 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, 8793e8800. Jin, Y., Chen, G., Fu, Z., 2015. Effects of TBEP on the induction of oxidative stress and endocrine disruption in TM3 Leydig cells. Environ. Toxicol. http://dx.doi.org/ 10.1002/tox.22137. 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 314e1, 76e83. Li, J., Yu, N., Zhang, B., Jin, L., Li, M., Hu, M., Zhang, X., Wei, S., Yu, H., 2014. Occurrence of organophosphate flame retardants in drinking water from China. Water Res. 54, 53e61. Liu, X., 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. 114e115, 173e181. Liu, X., 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. 134e135, 104e111. Ma, Y., Han, J., Guo, Y., Lam, P.K.S., Wu, R.S.S., Giesy, J.P., Zhang, X., Zhou, B., 2012. Disruption of endocrine function in in vitro H295R cell-based and in in vivo assay in zebrafish by 2,4-dichlorophenol. Aquat. Toxicol. 106e107, 173e181. Ma, Z., Tang, S., Su, G., Miao, Y., Liu, H., Xie, Y., Giesy, J.P., Saunders, D.M., Hecker, M., Yu, H., 2016. Effects of tris (2-butoxyethyl) phosphate (TBOEP) on endocrine axes during development of early life stages of zebrafish (Danio rerio). Chemosphere 144, 1920e1927. Marklund, A., Andersson, B., Haglund, P., 2005. Organophosphorus flame retardants and plasticizers in Swedish sewage treatment plants. Environ. Sci. Technol. 39,
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