Embryonic exposure to butachlor in zebrafish (Danio rerio): Endocrine disruption, developmental toxicity and immunotoxicity

Ecotoxicology and Environmental Safety 89 (2013) 189–195

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Ecotoxicology and Environmental Safety journal homepage: www.elsevier.com/locate/ecoenv

Embryonic exposure to butachlor in zebrafish (Danio rerio): Endocrine disruption, developmental toxicity and immunotoxicity Wenqing Tu a, Lili Niu b, Weiping Liu b, Chao Xu a,n a b

Research Center of Environmental Science, College of Biological and Environmental Engineering, Zhejiang University of Technology, Hangzhou 310032, PR China Institute of Environmental Science, College of Environmental and Resource Sciences, Zhejiang University, Hangzhou 310058, PR China

a r t i c l e i n f o

abstract

Article history: Received 13 July 2012 Received in revised form 5 November 2012 Accepted 29 November 2012 Available online 5 January 2013

Butachlor is a chloroacetanilide herbicide widely employed in weeding important crops. Recently, the study of the possible toxic effects of butachlor in non-target organisms has increased substantially. However, the endocrine disruption, developmental toxicity and immunotoxicity effects of butachlor in fish have not been fully investigated in previous studies. In the present study, zebrafish embryos were exposed to a range of butachlor concentrations from 4 to 20 mM to evaluate the embryonic toxicity of butachlor until 84 hours postfertilization (hpf). The results demonstrated that butachlor was highly toxic to zebrafish embryos, hindering the hatching process, resulting in a series of malformations and followed by mortality. The malformations observed included pericardial edema (PE) and yolk sac edema (YSE), which showed concentration-dependent responses. The analysis of endocrine gene transcription indicated that butachlor significantly induced the expression of the estrogen-responsive gene Vtg1 but had no effect on the expression of the ERa gene. The innate immune system appeared to be another possible target of butachlor. At 72 hpf, butachlor significantly up-regulated the innate immune system-related genes, including IL-1b, CC-chem, CXCL-C1c and IL-8. These data suggest that butachlor causes developmental toxicity, endocrine disruption and immune toxicity in the zebrafish embryo. Bidirectional interactions between the endocrine system and the immune system might be present, and further studies are needed to determine these possible pathways. & 2012 Elsevier Inc. All rights reserved.

Keywords: Butachlor Innate immune Developmental toxicity Estrogen-responsive Zebrafish

1. Introduction The pesticide contamination of aquatic ecosystems has become a concern worldwide. Many pesticides have been identified as actual or potential endocrine disruption chemicals (EDCs) in aquatic organisms. Chloroacetanilide herbicides are among the most frequently used herbicides. They play a vital role in increasing crop yield in agricultural production through their use as a preemergence herbicide in the control of undesirable grasses and broadleaved weeds (Singh and Pillsi, 1993). The results of the monitoring of ground water and of surface waters, such as lakes and rivers, for pesticides have consistently demonstrated the presence of chloroacetanilide herbicides (Kalkhoff et al., 1998). The active ingredients of chloroacetanilide herbicides include acetochlor, metolachlor, alachlor and butachlor. In Asian countries, butachlor is applied at extremely high levels, especially in rice paddy (Abdullah et al., 1997). Like most other chloroacetanilide herbicides, butachlor has been identified as a possible carcinogen and has high aquatic toxicity (Dearfield et al., 1999). The suggested

n

Corresponding author. E-mail address: [email protected] (C. Xu).

0147-6513/$ - see front matter & 2012 Elsevier Inc. All rights reserved. http://dx.doi.org/10.1016/j.ecoenv.2012.11.031

application concentration of butachlor for use in paddy fields is 10.7 mM to 150 mM (Dearfield et al., 1999; Chen et al., 2007) and the concentration detected in rivers range from 0.01 mg/L to 0.43 mg/L (Tsuda et al., 1997). However, even in this concentration range, acute toxicity tests have indicated that butachlor is extremely toxic to fish (Meng et al., 2007; Cagauan, 1995). However, aquatic toxicity assessments of butachlor at lower concentrations in fish species, especially in the early life stages, have only received limited attention. It is well known that endocrine disruptions can result in developmental malformations at early life stages. Interactions with hormone receptors represent one major mechanism of endocrine disruption. In teleost fish, certain EDCs can mimic endogenous estrogen to induce mRNA expression in estrogen receptors (ERs) and vitellogenins (Vtgs). Thus, the gene transcription of ERs and Vtgs have become frequently monitored biomarkers of estrogenic compounds in the aquatic environment (Arukwe et al., 1997; Kime et al., 1999; Jin et al., 2008). Previous reports have demonstrated that the transcription of two types of Vtgs (Vtg1, Vtg2) markedly increased in the liver of adult zebrafish after nonylphenol or 17b-estradiol (E2) treatments (Jin et al., 2009A). Many pesticides have also showed an ability to induce Vtgs in male adult and larval fish (Jin et al., 2008; Jin et al., 2009B).

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In addition, previous investigations have indicated bidirectional interactions involving the endocrine and immune systems in fish (Ahmed, 2000; Lutton and Callard, 2006). The components of the immune and reproductive systems overlap in fish, as they do in mammals. The recent literature suggests that immune parameters may be used as biomarkers of contamination by EDCs in teleost fish (Ahmed, 2000; Milla et al., 2011). Previous studies have demonstrated that EDCs could affect the components of the immune system, including leukocytes, macrophage activity and the innate and acquired immune systems (Milla et al., 2011). In early developmental stages, the innate immune system of fish is the only source of protection against infection by xenobiotics. As the gene transcription involved in innate immunity is most likely modulated by sex steroids, several of these genes have been demonstrated to be targets for estrogenic compounds (Williams et al., 2007). As a result, fish immunotoxicity may not only serve as an indicator of contamination but may also offer a way of understanding the mechanisms of endocrine disruption for certain chemicals. Zebrafish (Danio rerio) represent one of the most widely utilized and well-established models for the study of ontogenetic development and for toxiogenomic studies (Kimmel et al., 1995). The transparent embryos allow morphological observations of several lethal and nonlethal endpoints during the early life stages. Recently, many studies have confirmed that the exposure of zebrafish embryos to pesticides results in various nonlethal malformations, including yolk sac edema (YSE), pericardial edema (PE), crooked body and endocrine disruptions (Jin et al., 2009; Xu et al., 2008; Sun et al., 2010A). Zebrafish embryos also offer another important advantage as models for studies in immunotoxicology because they rely solely on the innate immune system for host defense for the first 30 days (Rojo et al., 2007). Sets of genes that are related to the innate immune system, such as IL1b, CC-chem, CXCL-C1c and IL-8, have been clearly characterized in zebrafish and greatly facilitate studies of toxicological mechanisms in this species (Rojo et al., 2007; Hermann and Kim 2005). In the present study, butachlor was evaluated for adverse effects in embryonic zebrafish, and its developmental toxicity and mortality were investigated. The expression of the estrogenresponsive genes vtg1 and ERa was also evaluated. To further understand the potential mechanisms of developmental toxicity and endocrine disruption in the early developmental stages of zebrafish, the study also investigated genes related to the immune system.

2. Materials and methods 2.1. Chemicals and materials Butachlor (98.5% purity) was purchased from Sigma–Aldrich (St. Louis, MO, USA). A stock solution of butachlor (2000 mg/L) was prepared in HPLC-grade nhexane and stored at 4 1C in darkness. The test solutions for the following experiments were prepared by diluting the stock solution in DMSO (analytical pure). The remaining chemicals used in this study were of analytical grade.

2.2. Zebrafish husbandry and embryo collection Zebrafish (Danio rerio) at the juvenile stage were obtained from the Institute of Hydrobiology of Chinese Academy of Science (Wuhan, China), cultured and reproduced in our laboratory. The zebrafish were maintained at a constant temperature (277 1 1C) on a light/dark cycle of 12 h: 12 h in a flow-through system with charcoal-filtered water (pH 7.2–7.6; hardness 44–61 mg/L CaCO3). The zebrafish were fed twice daily with live brine shrimp. Embryos were obtained from spawning adults in groups containing approximately 10 males and 5 females maintained in glass tanks. Spawning began in the morning when the light was turned on. The collection procedure followed that of Westerfield (1993).

2.3. Butachlor exposure experiments To investigate the developmental toxicity of butachlor during the embryonic stage of zebrafish, fertilized embryos were randomly distributed in 24-well plates containing different concentrations of butachlor (0, 4, 8, 12, 16, 20 mM) for 84 h. The procedural details were consistent with those specified by the OECD test guidelines (Braunbeck and Lammer, 2006). Briefly, newly fertilized embryos were first exposed to chambers with known concentrations of butachlor. Half an hour later, any non-fertilized embryos were removed, and the fertilized embryos were randomly placed in 24-well plates (Costar, Corning Inc, NY, USA) using a pipette. To eliminate interference, twenty wells of each 24-well plate were filled with 2 mL of the same concentration of butachlor in Hank’s solution (137 mM NaCl, 5.4 mM KCl, 0.25 mM Na2HPO4, 0.44 mM KH2PO4, 1.3 mM CaCl2, 1.0 mM MgSO4, and 4.2 mM NaHCO3) with 0.5% DMSO. The other four wells were used as internal controls and contained Hank’s solution with a solvent vehicle control (0.5% DMSO). Each 24-well plate was considered as one replicate for each concentration, and three replicates were used at each concentration. The plates were incubated at 287 1 1C with a photoperiod of 14 h light/10 h dark. The development of the embryos was monitored under an inverted dissecting microscope (Leica Micosystems, Wetzlar, Germany) until 84 hpf. Yolk sac edema, PE, hatching success and mortality were examined and recorded each 12 h. During exposure, dead embryos/larvae and detritus were promptly removed. Half of the volume of the exposure solutions was renewed daily. To investigate the influence of butachlor on the endocrine system and the innate immune system of the embryos, 15 newly fertilized embryos were randomly assigned to a 250 mL glass beaker containing 200 mL of the respective treatment solutions and were exposed under the same conditions as described above for 3 days. The exposure time and the concentration of selection is based on the results of the embryo developmental toxicity test of butachlor. Four replicates were run for each concentration, and the exposure solutions were renewed completely on a daily basis. At the end of exposure, at least 10 zebrafish larvae per treatment were pooled together as one sample for gene transcription analysis. The larvae were collection and placed on dry ice, and the samples were stored at  80 1C for subsequent analysis.

2.4. Analytical analysis of exposure solutions Samples of the exposure water for each treatment group were withdrawn from the experimental beakers shortly after exposure (T0) and before water renewal after 24 h (T24) of the experiments. The analysis of the concentrations of butachlor in water was performed with a high performance liquid chromatography (HPLC) system (Jasco, Japan). The samples were filtered through a 0.45-mm nylon filter (Millipore, MA) before injection. The injection volume was 20 mL. Chromatographic separation was performed with a C18 column (3.5 mm 150  4.6 mm, Waters, MA, USA). The mobile phase (acetonitrile:water¼ 72:25(V/V)) flow rate was 1 mL/min, the column temperature was 25 1C, and UV detection was performed at 215 nm. The detection limit for butachlor was 0.07 mg/L. The control solution was used as blanks for the method. To validate the integrity of the calibration curve, a calibration curve check standard was run before and after each batch.

2.5. RNA extraction, synthesis of cDNA and real-time PCR (quantitative PCR) The total RNA of each sample was extracted using TRIzol reagent (Takara Biochemical, Dalian, China) following the manufacturer’s instructions. The RNA quality in each sample was evaluated by 1% agarose gel electrophoresis and by the ratios of the 260:280 nm readings. The concentration of the RNA samples was estimated using the 260 nm reading. Reverse-transcription reactions to synthesize cDNA were performed with 500 ng of total RNA from each sample with a reverse transcriptase kit (Takara Biochemical, Dalian, China) according to the manufacturer’s protocols. The quantitation of IL-1b, CC-chem, CXCL-C1c, IL-8, Vtg1 and ERa using a 10 mL SYBR reaction mixture with specific primers was performed on the Mastercycler ep realplex (Eppendorf, Hamburg, Germany) according to the manufacturer’s instructions (SYBR Green Real-time PCR Master Mix, Takara Biochemicals, Dalian, China). The thermal cycling parameters were 95 1C for 1 min to activate the polymerase followed by 40 cycles of 95 1C for 15 s and 60 1C for 1 min. The mRNA levels were normalized to the corresponding b-actin value and calculated using the 2  DDCt method (Livak and Schmittgen, 2001). The primer sequences for IL-1b, CC-chem, CXCL-C1c, IL-8, Vtg1, ERa and a housekeeping gene (b-actin) are listed in Table 1.

2.6. Statistical analysis The data presented in this study were checked for normality and homogeneity of variances with Kolmogorov-Smirnov one-sample test and Levene test, respectively. The differences between the butachlor treatments and the corresponding controls were evaluated with a one-way analysis of variance (ANOVA followed by a Fisher post hoc test using the Stat View 5.0 program (SAS Institute Inc., Cary, NC, USA).

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Table 1 Sequences of primer pairs used in the real-time quantitative PCR reactions. Target gene

Primer sequences

Accession number

IL-1b IL-8 CXCL-Clc CC-chem Vtg1 ERa b-actin

F:50 -CATTTGCAGGCCGTCACA-30 R:50 -GGACATGCTGAAGCGCACTT-30 F:50 GTCGCTGCATTGAAACAGAA-30 R:5’-CTTAACCCATGGAGCAGAGG-3’ F:50 -CTGCTGCTTGCGGTAGTTTA-3’ R:50 -TCAACTTTGTCGCAGTTTGG-3’ F:50 -TGCAGCTCAACCAGAAGATG-30 R:50 CTTTGACGCATGGAGGATTT-30 F 50 -AACGAACAGCGAGAA AGAGATTG-30 R 50 GATGGGAACAGCGACAGGA-30 F 50 CCCACAGGACAAGAGGAAGA-30 R 50 CCTGGTCATGCAGAGACAGA-30 F 50 ATGGATGAGGAA ATCGCTGCC-30 R 50 CTCCCTGATGTCTGGGTCGTC-30

AY340959.1 XM_001342570.2 NM_001115060 BC162421.1 AB064320 D28954 AF057040

Table 2 Mean measured butachlor concentrations (mM 7 SE) in the water during the experiment. Butachlor concentration (mM)7 SE Nominal concentration Measured concentration T0 T24

4.00

6.00

8.00

12.00

16.00

20.00

3.43 7 0.58 2.88 7 0.05

5.34 70.26 5.65 70.17

6.81 7 2.03 6.55 7 2.42

11.07 7 2.04 10.92 7 1.70

14.97 75.23 14.32 70.14

20.617 1.56 19.23 77.03

The value p o 0.05 was considered statistically significant. The figures were drawn using Origin 8.0 (OriginLab, Northampton, MA, USA).

3. Results 3.1. Analytical quantification of exposure solutions The highest spiked butachlor concentration in the present study was below the known water solubility (Pal and Vanjara, 2001). Nominal solution concentrations were analytically quantified in triplicate. HPLC analysis of water samples showed that the exposure solutions ranged from approximately 72 to 103% of all the nominal concentrations (Table 2). 3.2. Zebrafish mortality following exposure to butachlor During the 84-h exposure of zebrafish embryos to butachlor, mortality, as identified by failure of the heartbeat and by somite formation with a non-detached tail (Jin et al., 2009; Shi et al., 2008), was monitored and recorded. The mortality of the embryos or larvae at four time points (48, 60, 72 and 84 hpf) is shown in Fig. 1. The results showed that mortality increased with increasing concentrations of butachlor and time of exposure. However, significant mortality was only observed at 84 hpf at concentrations of 16 mM and 20 mM. The mortalities were 93.3% and 100% for 16 mM and 20 mM exposure, respectively. No significant difference in mortality was recorded in the control group or at the other concentrations. The LC50 value of butachlor at 84 hpf was 14 mM. 3.3. Hatching success following exposure to butachlor In the present study, the hatching rates at 48, 60, 72 and 84 hpf were recorded to evaluate the influence of butachlor on the hatching process. The hatching rate is defined as the number of hatched embryos at specific exposure time is divided by that of all embryos at the beginning. As shown in Fig. 2, hatching began after 48 hpf. The hatching rate increased daily (48 hpf to 84 hpf) in the control, 4 mM and 8 mM groups. The percentage of hatched embryos decreased significantly from 93.3% to 73.3% and 50.0% for the 0, 4 and 8 mM groups at 84 hpf, respectively. The decrease of hatching rate from 64 to 82 hpf at the exposure concentration

Fig. 1. Mortality of zebrafish embryos exposed to various concentrations of butachlor for different periods. The error bars represent the standard deviation (SD) of the mean. The asterisks indicate significant differences from the control group * at po 0.05 and ** at po 0.01.

of 16 mM mainly due to the increasing of mortality from 64 to 82 hpf at 16 mM (see Fig. 1). Note that a slight acceleration in hatching relative to the control occurred by 60 hpf at a concentration of 8 mM, as found in a previous study (Jin et al. 2009). In comparison, the hatching process was significantly hindered by butachlor at higher concentrations (e.g., 12, 16 and 20 mM). For example, only a few embryos hatched at 12 mM, and no hatched zebrafish larvae were observed at either 16 mM or 20 mM. 3.4. Malformations of zebrafish induced by butachlor Exposure to butachlor resulted in a series of developmental malformations, including YSE, PE, crooked body and coagulation. The percentage of PE or YSE is defined as the number of PE or YSE embryos is divided by that of all embryos at the beginning. In the present study, PE and YSE were recorded in detail from 48 to 84 hpf. Fig. 3 shows that no malformations were observed prior to 48 hpf. At 48 hpf, YSE was the major malformation. From 60 to 84 hpf, more developmental malformations appeared, primarily

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was also observed in PE, but it occurred at a later stage of hatching. At 48 hpf, no PE was observed even at the highest concentration tested. At 60 hpf, PE was observed even at the lowest concentration. 3.5. Estrogen-responsive gene transcription

Fig. 2. Hatching rates of zebrafish embryos exposed to various concentrations of butachlor for different periods. The error bars represent the standard deviation (SD) of the mean. The asterisks indicate significant differences from the control group * at p o0.05 and ** at p o0.01.

As is shown in Fig. 1, mortality at concentrations of 16 and 20 mM at 84 hpf were 93.3% and 100%, respectively. While at 72 hpf, the mortality is relatively high when the concentration is higher than 12 mM. Therefore, the exposure time for evaluation of gene transcription lasted only 72 h and the highest concentration of exposure was 12 mM. The expression of Vtg1 and ERa mRNA was measured in the larval zebrafish after exposure to various concentrations of butachlor for 3 days (Fig. 4). Compared with the control, the Vtg1 transcripts were slightly up-regulated at lower concentrations of butachlor, but a significant increase occurred at 8 and 12 mM. For example, the induction levels of Vtg1 were 5.39 and 4.44-fold compared with the control at concentrations of 8 mM and 12 mM, respectively. However, no significant induction for ERa was observed at any concentration tested. Similar results have been reported previously (Sun et al., 2010B). 3.6. Expression of genes related to the innate immune system The expression of IL-1b, CC-chem, CXCL-C1c and IL-8 mRNA was determined with quantitative real-time PCR after 3 days exposure to butachlor (Fig. 5). The results indicated that the transcription levels of IL-1b and IL-8 in the butachlor-exposed zebrafish were significantly increased at 12 mM, with 3.95- and 4.47-fold induction, respectively, relative to the control (po0.05) (Fig. 5A and D), whereas no significant differences were observed in the CXCL-Clc mRNA level compared with the control for all treatments (Fig. 5C). Moreover, the expression of CC-chem was significantly up-regulated at 8 mM butachlor exposure. (Fig. 5B). However, the mRNA expression of these genes showed a slight decrease at the highest concentration.

4. Discussion

Fig. 3. Morphological abnormalities in zebrafish exposed to various concentrations of butachlor following 84-h exposure. (A) pericardial edema; (B) yolk sac edema. The error bars represent the standard deviation (SD) of the mean.

PE and YSE, which showed dose-dependent relationships. For example, the percentage of embryos showing YSE ranged from 10% in 8 mM to 73.3% in 20 mM at 48 hpf. By 84 hpf, due to the increasing mortality, the percentage of YSE was significantly lower at higher concentrations (16 and 20 mM). A similar pattern

The objective of the present study was to perform an examination of the aquatic toxicity of butachlor in zebrafish embryos. The results demonstrate that butachlor exposure during early development can cause lethal and non-lethal developmental malformations, induce the expression of the estrogenresponsive gene Vtg1 and has a substantial impact on the innate immune system. The sensitivity of embryos to butachlor exposure resulted in significant mortality in a concentration-dependent manner, especially at higher concentrations after 84-h exposure. The LC50 (96 h) values of butachlor to fish range from 0.14 to 0.52 mg/L, concentrations highly toxic to fish (Tomlin, 1994). In the present study of embryos, the LC50 value was 14 mM (equivalent to 4.4 mg/L), higher than the result for adult fish in the Tomlin et al. study. The chorion of the embryo might act as a barrier for pesticides and thus result in a higher LC50 value. A similar phenomenon has also been observed with lambda-cyhalothrin in zebrafish embryos (Xu et al., 2008). However, even the highest concentration in our study is still much lower than the recommended concentration of butachlor for use in paddy fields (up to 150 mM). This observation indicates that the use of butachlor in paddy fields could cause serious ecotoxicity. Exposure to butachlor not only delayed the hatching time but also reduced the hatching rate. For the zebrafish embryos exposed to the control and low concentrations (i.e., 4 mM and 8 mM),

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8

8

**

4

6 0 12

2 0

*

8 6

*

Relative mRNA levels

Relative mRNA levels

4

*

8 4 0 16 12 8 4 0 12

4

8

2 0

*

4 0

0

4

6

8

12

Concentration of Butachlor (µM) Fig. 4. Expression of (A) ERa and (B) Vtg1 in larval zebrafish exposed to various concentrations of butachlor for 72 h. Values were normalized against b-actin (a housekeeping gene) and the mean mRNA expression value 7SD (n¼4) relative to the values in the controls. The error bars represent the standard deviation (SD) of the mean. The asterisks indicate significant differences from the control group * at po 0.05 and **at p o 0.01.

hatching began after 48 hpf, whereas no larvae hatched at the highest concentration (20 mM) even at 84 hpf. This phenomenon indicates that the hatching process was inhibited by butachlor at higher concentrations (e.g., 20 mM). It is interesting to note that exposure to 8 mM accelerated the hatching process compared to the control and 4 mM at 60 hpf and 72 hpf, respectively. This outcome might be a result of hormesis (Kaiser, 2003). Hatching time represents an important developmental marker for larval aquatic organisms. Xenobiotics may alter the signals in the biochemical and developmental pathways needed for the embryo to free itself from the chorion (Sano et al., 2008; Nechaev and Pavlov, 2004). These results suggest that butachlor can affect the regulation of hatching in the developmental process of the zebrafish embryo and that the effect depends on the concentration of butachlor. Developmental malformations during embryonic development are commonly observed after exposure to xenobiotic compounds (Hill et al., 2005; Carney et al., 2006). Exposure to butachlor resulted in malformations, such as pericardial and yolk edema, during embryonic development. In embryos, developmental toxicity is more sensitive than acute toxicity and offers more endpoints. After 48 hpf, more developmental abnormalities appeared and increased in a concentration-dependent manner. At 72 hpf, the PE and YSE at 12 mM were 86.7% and 96.7%, respectively. These values were much higher than those of the control group. The results showed that butachlor had a significant developmental toxicity in zebrafish in early development.

0

4 6 8 Concentration of Butachlor (µM)

12

Fig. 5. Expression of (A) IL-1b, (B) CC-chem, (C) CXCL-Clc and (D) IL-8 in larval zebrafish exposed to various concentrations of butachlor for 72 h. Values were normalized against b-actin (a housekeeping gene), and the mean mRNA expression value7 SD (n¼ 4) relative to the values in the controls. The error bars represent the standard deviation (SD) of the mean. The asterisks indicate significant differences from the control group * at p o 0.05 and ** at p o 0.01.

An increasing number of studies have shown that environmental toxicants, including pesticides and industrial solvents, produce deleterious effects on the development of non-target organisms by disrupting the endocrine system. In the present study, the gene transcription of Vtg1 and ERa was studied to better understand the mechanisms by which butachlor affects the development of the zebrafish embryo. The results obtained from our study show that exposure to butachlor caused the induction of Vtg1 gene transcription in larval zebrafish, especially at higher concentrations. However, ERa gene transcription showed no significant difference between butachlor exposure and the control. The reason for this difference in gene transcription is that Vtg genes are usually more sensitive to estrogen exposure than the ERa gene. The mechanism of action of natural estrogens, such as E2, is that they first combine with a specific ER. The estrogen–ER then enters a complex interaction with estrogen-responsive elements (EREs), which are the target promoter genes that regulate gene transcription (Driscoll et al., 1998; Klinge, 2001). Vitellogenin, the egg yolk protein precursor, is generated in the liver in response to stimulation by estrogenic compounds, transported to the ovary through the blood stream, and incorporated into the oocytes in teleost fish (Wallace, 1985). Generally, the estrogen-responsive Vtg genes are produced only by the liver in adult female fish and are inactive in male and sexually immature fish (Brion et al., 2002). However, a number of previous studies have shown that Vtg synthesis can be induced in adult male or sexually immature fish after exposure to certain EDCs, such as alkylphenolic compounds, phytoestrogens, synthetic estrogens

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and pesticides (Driscoll et al., 1998). Although the information presented here regarding the estrogenic activities of butachlor on larval zebrafish is limited, the results indicated that butachlor can act as a potential estrogenic chemical, significantly inducing the expression of the estrogen-responsive gene Vtg1. This result is consistent with previous reports that butachlor can adversely affect the endocrine system in adult zebrafish (Chang et al., 2011). A number of studies have shown that sex steroids and related EDCs can affect all aspects of the immune system in teleosts and vice versa. Environmentally realistic levels of exposure to contaminants that have effects on the immune system can cause endocrine disruption (Milla et al., 2011). The innate immune system is the sole defense against infection by pathogens in the early developmental stage of zebrafish (Trede et al., 2004). The adaptive immune system is inactive during the first weeks of zebrafish development (Nayak et al., 2007). Therefore, the early developmental stage of zebrafish becomes a useful model for investigating innate immune function. More recently, numerous reports have shown that the mRNA expression of genes related to the innate immune system was induced or suppressed by bacteria, viruses or environmental chemicals (Phelan et al., 2004; Eder et al., 2008). Therefore, research on the innate immune system function of zebrafish helps to achieve a full understanding of the toxicity of butachlor. The present study showed that the mRNA level of the innate immune-related genes is significantly induced by butachlor in embryonic zebrafish. The expression of the IL-lb, CXCL-C1c, CC-chem and IL-8 genes in early developmental stages of zebrafish was induced rapidly and effectively in a concentration-dependent manner after exposure to butachlor for 3 days. An effectively functioning immune system plays a vital role in fish in response to environmental changes. Once the system is disturbed, the cellular innate immune responses may fail to defend against infection (Magnadottir, 2006; Whyte, 2007). The IL-1b gene plays a vital role in recruiting phagocytes to the site of infection. It activates neutrophils and macrophages and stimulates their recruitment to the site of the injury. The other genes, including IL-8, CXCL-C1c and CC-chem, show the ability to attract and activate leukocytes and to influence their potential roles as mediators of inflammation. Thus, butachlor can potentially induce an immune response in the early development of zebrafish. It had reported that administration of high levels of estrogen can induce thymic involution (Zoller and Kersh, 2006). On the other side, pathologies in ovaries, testes and thyroid endocrine tissues could be induced by removal of the thymus (Hattori and Brandon, 1979; Tung et al., 1987). Therefore, it may be inferred that the communication between endocrine disrupters and immune system was mediated by hormones and cytokines (Ahmed, 2000; Lutton and Callard, 2006). At present, because the interactions between the endocrine and immune systems produce the same variation in the tendency to respond to butachlor, the impact on the defense against infection by butachlor, operating through the innate immune system, is a possible explanation for the observation that butachlor significantly induced the expression of Vtg1 during the early development of zebrafish. The study of bidirectional interactions between the endocrine system and the immune system can help to produce a better understanding of the mechanism underlying the toxic effects of butachlor.

5. Conclusion Developmental toxicity, endocrine disruption and innate immune toxicity were evaluated for butachlor with embryonic zebrafish. Butachlor caused significant mortality, inhibited embryo hatching and resulted in a number of morphological

abnormalities, primarily YSE and PE. Moreover, the effects on the transcription of the estrogen-responsive gene Vtg1 increased with increasing concentrations of butachlor. This gene was more sensitive than ERa to changes in the concentration of butachlor. Furthermore, the expression of genes that are closely related to the innate immune system, including IL-1b, CC-chem, CXCL-C1c and IL-8, was also significantly affected. Further testing is necessary to better understand the mechanisms underlying the developmental toxicity of butachlor and the bidirectional interaction between the endocrine and immune systems.

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