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Keywords:

  • Pentachlorophenol;
  • Thyroid hormone;
  • Hypothalamic–pituitary–thyroid axis;
  • Gene expression;
  • Zebrafish

Abstract

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. Acknowledgment
  8. REFERENCES

Pentachlorophenol (PCP) is frequently detected in the aquatic environment and has been implicated as an endocrine disruptor in fish. In the present study, 4-month-old zebrafish (Danio rerio) were exposed to 1 of 4 concentrations of PCP (0.1, 1, 9, and 27 µg/L) for 70 d. The effects of PCP exposure on plasma thyroid hormone levels, and the expression levels of selected genes, were measured in the brain and liver. The PCP exposure at 27 µg/L resulted in elevated plasma thyroxine concentrations in male and female zebrafish and depressed 3, 5, 3'-triiodothyronine concentrations in males only. In both sexes, PCP exposure resulted in decreased messenger RNA (mRNA) expression levels of thyroid-stimulating hormone β-subunit (tshβ) and thyroid hormone receptor β (trβ) in the brain, as well as increased liver levels of uridine diphosphoglucuronosyl transferase (ugt1ab) and decreased deiodinase 1 (dio1). The authors also identified several sex-specific effects of PCP exposure, including changes in mRNA levels for deiodinase 2 (dio2), cytosolic sulfotransferase (sult1 st5), and transthyretin (ttr) genes in the liver. Environmental PCP exposure also caused an increased malformation rate in offspring that received maternal exposure to PCP. The present study demonstrates that chronic exposure to environmental levels of PCP alters plasma thyroid hormone levels, as well as the expression of genes associated with thyroid hormone signaling and metabolism in the hypothalamic-pituitary-thyroid (HPT) axis and liver, resulting in abnormal zebrafish development. Environ Toxicol Chem 2014;33:170–176. © 2013 SETAC


INTRODUCTION

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. Acknowledgment
  8. REFERENCES

Pentachlorophenol (PCP) is a broad-spectrum biocide that has been widely used in wood preservatives, pesticides, and disinfectants [1]. Because of its toxicity, PCP has been restricted or banned in many countries since the 1980s; however, it is still used as a wood preservative [2]. In 2002, approximately 11 million pounds of PCP were produced in the United States [3], and in China, the annual production of PCP increased to 3000 tons in 2003 after the re-emergence of schistosomiasis in several provinces [4]. The continued use of PCP throughout the world means it is still likely to result in environmental contamination, with potential risks to human health and the environment.

Environmental monitoring has demonstrated that PCP exists in various environmental media (e.g., water, soil, sediment, and aquatic organisms) and human samples [5-7], and it is a particularly common contaminant in water [8, 9]. For instance, the PCP level of surface waters is typically in the range of 0.1 µg/L to 1.0 µg/L, whereas an elevated PCP concentration can be found in groundwater (3–23 µg/L) and surface water (0.07–31.9 µg/L) in the United States, where the treatment of wood products with PCP has been conducted [6]. In China, PCP contamination in the Yangtze River was most severe among 7 of its watersheds and 3 drainage areas, with a median value of 0.06 µg/L (up to 0.59 µg/L) [8], and greater concentrations of PCP (up to 103.7 µg/L) were detected in Dongting Lake [10]. High concentrations of PCP in water may lead to a high risk of exposure and negative biological effects in aquatic organisms as well as humans.

The impact of environmental chemicals on the thyroid endocrine system has received much attention in recent years [11]. Endocrine system disrupting chemicals can have a direct impact on thyroid hormone synthesis, transport, binding, catabolism, and clearance of circulating thyroid hormones. Limited information shows that PCP might disrupt thyroid endocrine functions. Epidemiological studies have reported that exposure to low levels of PCP in the environment is associated with decreased thyroid hormone levels in human neonates [12]. A link between PCP exposure and disrupted thyroid hormone signaling has been experimentally demonstrated in rats [13-15], ewes [16], and Fischer rat thyroid cell line FRTL-5 [17]. For instance, a pronounced decrease in circulating thyroxine (T4) and 3, 5, 3'-triiodothyronine (T3) levels were observed in female rats treated with 3 mg PCP/kg and 30 mg PCP/kg body weight for 28 d [16]. Similarly, oral administration of PCP during development significantly increased the messenger RNA (mRNA) expression of thyroid hormone receptor β1 and synapsin I in the brain of rats, and reduced plasma levels of total T4 (TT4) in dams and 3-wk-old pups [14]. Treatment with pentachlorophenate sodium (PCP-Na) significantly affected the thyroid endocrine system and altered mRNA expression of thyroid hormone receptors and deiodinases in the rat liver. In fish, PCP exposure decreases serum T4 levels [18].

The thyroid hormones play a crucial role in the regulation of development, growth, immunity, metabolism, reproduction, and behavior in vertebrates [19]. In fish, thyroid homeostasis is controlled primarily by the hypothalamic-pituitary-thyroid (HPT) axis [20]. Recently, zebrafish (Danio rerio) thyroid system has become a very popular vertebrate model to screen for thyroid-disrupting chemical pollutants [18, 21-23]. Several studies have suggested that PCP exposure may affect the expression level of genes associated with thyroid hormone metabolism and signaling in rats [14, 15] and the HPT axis in zebrafish larvae [18]; however, the thyroid hormone endocrine disruption and environmental risk of lower concentration of PCP for adult fish remains unclear. As such, in the present study we have evaluated the effects of lower concentration of PCP exposure on thyroid hormone homeostasis in zebrafish, using plasma thyroid hormone levels and gene expression involved in the HPT axis, to gain an improved understanding of PCP's effect on thyroid hormone function.

MATERIALS AND METHODS

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. Acknowledgment
  8. REFERENCES

Chemicals

Pentachlorophenol (purity >99%) was purchased from Dr. Ehrenstorfer GmbH. It was dissolved in dimethylsulfoxide and stored at 4 °C. The TRIzol reagent and PrimeScript Reverse Transcription Reagent kit were purchased from TaKaRa, and the SYBR Real-time polymerase chain reaction Master Mix was purchased from Tiangen. Enzyme-linked immunosorbent assay kits for T3 and T4 were purchased from EIAab Science. All other chemicals used in the present study were of analytical grade.

Zebrafish maintenance and experimental design

Adult zebrafish (aged 4 mo; AB strain; males: 302 mg ± 24 mg; females: 423 mg ± 36 mg) were raised in 12-L glass tanks, which contained 10 L dechlorinated tap water. Each tank contained 10 randomly selected zebrafish (5 females and 5 males). Fish were subjected to a 14:10 h light:dark at a constant temperature of 28 ± 0.5 °C, according to previously described methods [22]. Before exposure to PCP, zebrafish were acclimated in tanks for 1 wk. Fish were exposed to different doses of PCP at concentrations of 0 µg/L, 0.1 µg/L, 1 µg/L, or 27 µg/L for 70 d. These concentrations were selected based on the provisional guideline concentrations of PCP in drinking water recommended by the World Health Organization in 2003 (9 µg/L) and the US Environmental Protection Agency in 1997 (1 µg/L) [9]. Throughout the experimental period, the water in each tank was replaced daily with fresh solutions at the appropriate concentration of PCP. Both the control and exposure groups received 0.001% (vol/vol) dimethylsulfoxide. Three replicate tanks were available for the control and PCP-exposed groups. All animals were treated humanely and with the aim of alleviating any suffering. They were maintained in accordance with guidelines for the care and use of laboratory animals of the National Institute for Food and Drug Control of China.

Thyroid hormone assays

After 70 d exposure, the adult fish were anesthetized in 0.03% MS-222. The body weight was recorded, and blood was collected from the caudal vein of each fish. The blood samples from 5 fish of the same sex were pooled as 1 replicate (∼40 µL). The pooled blood was then centrifuged at 7000 g for 5 min at 4 °C, and the plasma was collected and stored at −80 °C until analysis.

Plasma total T4 (TT4) and T3 (TT3) levels were measured in adult zebrafish as described previously [22]. In brief, plasma thyroid hormone levels were measured using enzyme-linked immunosorbent assay test kits (Uscnlife), following the manufacturer's instructions. The detection limits, intra-assay and inter-assay variations reported by the manufacturer are 1.2 ng/mL, 4.3%, and 7.5% for TT4, and 0.1 ng/mL, 4.5%, and 7.2% for TT3, respectively.

Quantitative real-time PCR

The liver and brain (including hypothalamus and pituitary) were dissected and preserved in TRIzol reagent (TaKaRa) for RNA sample preparation. Extraction, purification, and quantification of total RNA, first-strand complementary DNA synthesis, and quantitative real-time polymerase chain reaction assays were carried out as previously described [23]. Briefly, total RNA was isolated using TRIzol regent, and digested with RNase-free DNaseI (Promega), following the manufacturer's instructions. The concentration of total RNA was assayed at 260 nm and 280 nm using a spectrophotometer (M2, Molecular Devices), and the purity of RNA in each sample was verified by determining the A260/A280 ratio and confirmed by agarose-formaldehyde gel electrophoresis with ethidium bromide staining. The purified RNA was used immediately for reverse transcription or stored at −80 °C until analysis.

Synthesis of first-strand complementary DNA was performed using a PrimeScript Reverse Transcription Reagent Kit (TaKaRa) following the manufacturer's instructions. The quantitative real-time polymerase chain reaction was performed using a SYBR Green PCR kit (Tiangen) on an Agilent Mx3005P qPCR System (Agilent Technologies). The primer sequences of the selected genes were obtained using the online Primer 3 program [24] and are shown in Table 1. Genes responsible/involvement for thyroid hormone synthesis (tshβ), transport (ttr), binding (trβ), and metabolism (deiodinase 1 [dio1], deiodinase 2 [dio2], uridine diphosphoglucuronosyl transferase [ugt1ab], and cytosolic sulfotransferase [sult1 st5]) along the HPT axis and liver were selected. The thermal cycle was set at 95 °C for 2 min, followed by 40 cycles at 95 °C for 15 s, 60 °C for 15 s, and 72 °C for 1 min, and a final cycle of 95 °C for 15 s, 60 °C for 1 min, and 95 °C for 15 s. Gene expression was measured in quadruplicate and repeated 3 times. The housekeeping gene ribosomal protein L8 (rpl8) did not vary on chemical exposure (data not shown) and was used as an internal control. The mRNA expression of each gene was normalized to its corresponding rpl8 mRNA level. The relative mRNA expression was determined by the 2−ΔΔCt method.

Table 1. Primer sequences used in the present study
GeneSequence of the primer (5′-3′)Genbank accession number
  1. rpl8 = ribosomal protein L8; tshβ = thyroid-stimulating hormone β-subunit; trβ = thyroid hormone receptor β; ttr = transthyretin; dio1 = deiodinase 1; dio2 = deiodinase 2; ugt1ab = uridine diphosphoglucuronosyl transferase; sult1 st5 = cytosolic sulfotransferase.

rpl8Forward: ttgttggtgttgttgctggtNM_200713
 Reverse: ggatgctcaacagggttcat 
tshβForward: gcagatcctcacttcacctaccAY135147
 Reverse: gcacaggtttggagcatctca 
trβForward: tgggagatgatacgggttgtNM_131340
 Reverse: ataggtgccgatccaatgtc 
ttrForward: cgggtggagtttgacactttBC081488
 Reverse: gctcagaaggagagccagta 
dio1Forward: gttcaaacagcttgtcaaggactBC076008
 Reverse: agcaagcctctcctccaagtt 
dio2Forward: gcataggcagtcgctcatttNM_212789
 Reverse: tgtggtctctcatccaacca 
ugt1abForward: ccaccaagtctttccgtgttNM_213422
 Reverse: gcagtccttcacaggctttc 
sult1 st5Forward: gaaagaggaccctgctcgtgNM_001199903
 Reverse: tttgccatggggttttctcg 

F1 generation endpoints

On the last day of exposure, 100 randomly selected eggs from each tank were separately cultured in glass dishes that contained 250 mL fresh water without PCP exposure until 4 d postfertilization. The number of larvae that exhibited malformations, the hatching rate, and the survival rate were determined at 4 d postfertilization.

Quantification of PCP in exposure solutions

Exposure solutions were renewed daily and sampled twice for measurement of PCP. Concentrations of PCP in exposure solutions were determined at 0 h (freshly prepared solutions) and 24 h (before renewal of exposure solutions) in both control and treatment groups. Quantification of PCP was performed using a published protocol [25]. Briefly, 500 mL of each sample was combined with 1 mL 6 mol/L hydrochloric acid, acidified to pH < 2, and then filtered through 0.45 µm nylon membrane filters (Jiuding High-Tech Filtration). The filtered samples were loaded on hydrophilic-lipophilic balanced cartridges preconditioned by methanol and water. After loading, the cartridge was rinsed with water and eluted with 5 mL ethyl acetate and 5 mL methanol. The eluate was reduced to near dryness (<0.1 mL) under a gentle stream of nitrogen, reconstituted to 0.5 mL with methanol. The samples were then immediately used for gas chromatography-mass spectrometry analysis. The calibration and quantification were performed by using an internal standard method with PCP standard prepared in ethyl acetate, with 2, 4, 6-tribromophenol serving as an alternative internal standard.

Quantification of PCP was performed using a gas chromatograph (Agilent 6890A) equipped with a mass-selective detector (Agilent 5975C) using electron impact mode. The gas chromatography column used for quantification was a capillary HP-5MS (5% phenyl/95% methylpoly siloxane, 30 m, 0.25 mm inner diameter, 0.1 µm film thickness, Agilent). Procedural blanks were analyzed simultaneously with every batch of 3 samples to check for interference or contamination from solvents or glassware. Recovery of 2, 4, 6- tribromophenol ranged between 85% and 98%. The detection limit was calculated as 3 times the procedural blank (0.05 µg/L for PCP).

Statistical analysis

All data were expressed as means ± standard error. The normality of the data was verified using the Kolmogorov-Smirnov test. The homogeneity of variances was analyzed by Levene's test. The differences between the control and each exposure group were evaluated by one-way analysis of variance followed by Tukey's post-test using SPSS 13.0 (SPSS). A p value < 0.05 was considered statistically significant.

RESULTS

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. Acknowledgment
  8. REFERENCES

Body condition in the F0 generation

No mortality was observed in the control and PCP-exposed groups, and no effects on the hepatic-somatic index and brain-somatic index were seen in either male or female fish (Table 2). The gonadal-somatic index was significantly decreased in females exposed to 27 µg/L PCP, but it was not significantly different in males or controls (Table 2). In males, 27 µg/L PCP exposure significantly inhibited body weight (0.29 g ± 0.01 g) compared with control (0.31 g ± 0.01 g).

Table 2. Somatic indices in zebrafish after exposure to pentachlorophenol (0 µg/L, 0.1 µg/L, 1 µg/L, 9 µg/L, 27 µg/L) for 70 da
PCP (μg/L)FemaleMale
00.1192700.11927
  • a

    The values represent mean ± standard error of 3 replicate tanks (5 females and 5 males for each replicate tank).

  • b

    Condition factor (K-factor): calculated as (wt/length3) × 100.

  • c

    BSI = brain weight × 100/body weight.

  • d

    HSI = liver weight × 100/body weight.

  • e

    GSI = gonad weight × 100/body weight.

  • *

    p < 0.05 indicates significant differences between treatment groups and the corresponding control group.

  • PCP = pentachlorophenol; BSI = brain size index; HSI = hepatosomatic index; GSI = gonadosomatic index.

Length (cm)3.50 ± 0.063.52 ± 0.053.56 ± 0.063.53 ± 0.053.65 ± 0.073.47 ± 0.033.47 ± 0.053.37 ± 0.043.49 ± 0.053.34 ± 0.04
Weight (g)0.44 ± 0.020.47 ± 0.030.44 ± 0.020.41 ± 0.020.44 ± 0.020.31 ± 0.010.34 ± 0.010.30 ± 0.010.33 ± 0.010.29 ± 0.01*
K-factorb1.02 ± 0.031.07 ± 0.030.97 ± 0.030.94 ± 0.030.90 ± 0.020.75 ± 0.020.82 ± 0.030.90 ± 0.010.77 ± 0.020.78 ± 0.02
BSI (%)c1.09 ± 0.051.20 ± 0.061.31 ± 0.091.19 ± 0.081.29 ± 0.051.77 ± 0.091.80 ± 0.111.86 ± 0.071.82 ± 0.071.91 ± 0.11
HSI (%)d1.72 ± 0.131.84 ± 0.112.19 ± 0.132.22 ± 0.122.02 ± 0.171.02 ± 0.470.84 ± 0.141.23 ± 0.131.29 ± 0.231.08 ± 0.11
GSI (%)e14.9 ± 1.5815.2 ± 1.1311.33 ± 1.2311.33 ± 1.2312.60 ± 0.83*1.21 ± 0.131.04 ± 0.131.23 ± 0.060.95 ± 0.131.15 ± 0.09

Plasma concentration of thyroid hormone

Exposure to PCP affected the plasma concentration of TT4 in both female and male zebrafish, while causing effects on TT3 concentrations only in males. In females, the plasma TT4 level showed a significant increase (44.7%) in the highest exposure group (Figure 1A). In males, the plasma TT4 level was significantly greater than that of controls by 161.5%, when exposed to 27 μg/L PCP (Figure 1A). In females, TT3 levels were not significantly changed after PCP exposure (Figure 1B); however, plasma TT3 levels in male zebrafish were significantly decreased by 34.5% and 38.3% when exposed to 9 and 27 μg/L PCP, respectively, compared with control levels (Figure 1B).

image

Figure 1. Thyroxine (T4) (A) and triiodothyronine (T3) (B) levels in zebrafish after exposure to 0.1, 1, 9, and 27 µg/L pentachlorophenol. The thyroid hormone levels in the zebrafish were expressed as ng/mL. All data are expressed as means ± standard error of the mean of 3 replicates (5 fish per replicate). Values that are significantly different from the control are indicated by asterisks. *p < 0.05; **p < 0.01.

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mRNA expression profile of selected genes

Several genes involved in the regulation, transport, binding, and metabolism of thyroid hormones were examined. In the brain of female zebrafish, tshβ gene expression was significantly downregulated by 47.7% in the 27 μg/L PCP-exposed group (Figure 2A). Similarly, in the males, a concentration-dependent downregulation of tshβ expression was observed, such that increasing concentrations of PCP were associated with incrementally lower tshβ expression (Figure 2B). Furthermore, exposure of males to 27 µg/L PCP also caused a significant downregulation of thyroid hormone receptorβ (trβ) (3.63-fold; Figure 2B).

image

Figure 2. The gene expression of tshβ and trβ in the brain of the (A) females and (B) males in the F0 adult zebrafish after exposure to 0.1 µg/L, 1 µg/L, 9 µg/L, and 27 µg/L pentachlorophenol. All data are expressed as means ± standard error of the mean of 3 replicates. Values that are significantly different from the control are indicated by asterisks. *p < 0.05; **p < 0.01.

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In the liver, dio1, dio2, ugt1ab, sult1st5, and ttr expression were examined. In females, dio1 expression was significantly decreased in a concentration-dependent manner, whereas a significant upregulation occurred in dio2 expression that was also dose-dependent (Figure 3A). The ttr gene expression was significantly upregulated by 2.19-fold in the 27 μg/L PCP-exposed group, and there was a strong dose-dependent increase in hepatic ugt1ab expression with an approximately 24-fold increase in expression observed in the group exposed to 27 μg/L PCP (Figure 3A). Conversely, sult1 st5 expression was significantly downregulated by PCP exposure at all concentrations studied (Figure 3A). In males, the expression of dio1 and dio2 were both significantly downregulated by PCP exposure at concentrations of 1 μg/L, 9 μg/L, and 27 μg/L (Figure 3B). In contrast to female fish, ttr expression was significantly downregulated by PCP in a concentration-dependent manner in the liver of male fish. The expression of ugt1ab was significantly increased in a concentration-dependent manner, with significant increases observed at doses of 1 μg/L, 9 μg/L, and 27 μg/L PCP (Figure 3B). Furthermore, treatment with the highest concentration of PCP (27 μg/L) significantly upregulated the expression of sult1 st5 (Figure 3B).

image

Figure 3. The gene expression of dio1, dio2, ugt1ab, sult1st5, and ttr in the liver of the (A) females and (B) males in the F0 adult zebrafish after exposure to 0.1 µg/L, 1 µg/L, 9 µg/L, and 27 µg/L pentachlorophenol. All data are expressed as means ± standard error of the mean of 3 replicates. Values that are significantly different from the control are indicated by asterisks. *p < 0.05; **p < 0.01.

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F1 generation toxicological endpoints

In the offspring of PCP-exposed fish, 76.3% of the control embryos hatched successfully at 4 d postfertilization, and no significant difference was seen in the hatching rate observed in any of the PCP-treated groups (Table 3). The rate of malformation (spinal curvature) was significantly increased to 1.7% and 1.8% after parental exposure to 9 μg/L and 27 μg/L PCP, respectively, compared with control (Table 3), whereas the recorded survival rates of embryos showed no significant differences after 70 d parental exposure to PCP (Table 3).

Table 3. Effects of pentachlorophenol on F1 zebrafish embryo/larvae (4 d postfertilization) developmental parameters
F1 larvae0 µg/L0.1 µg/L1 µg/L9 µg/L27 µg/L
  • a

    Values represent the mean ± standard error of 3 replicate tanks (100 eggs per tank).

  • b

    Values represent the mean ± standard error of 3 replicate tanks (15 larvae per tank).

  • *

    p < 0.05 indicates significant difference between treatment groups and the corresponding control group.

  • **

    p < 0.01 indicates significant difference between treatment groups and the corresponding control group.

Hatching rate (%)a76.33 ± 1.0467.00 ± 3.4269.83 ± 2.2764.00 ± 4.9082.67 ± 4.78
Surviving rate (%)a86.22 ± 2.3082.56 ± 3.7782.73 ± 2.8975.50 ± 3.1588.22 ± 0.44
Malformation rate (%)b0.83 ± 0.501.56 ± 0.201.83 ± 0.172.50 ± 0.17*2.67 ± 0.38**

Quantification of PCP in exposure solutions

No significant change in the water concentration of PCP was observed 24 h after use in comparison with freshly prepared solutions, although a decreasing trend was observed at each concentration. Mean concentrations of PCP 0 h (freshly prepared solutions) were 0.10 μg/L, 1.10 μg/L, 9.23 μg/L, and 33.64 μg/L in the 0.1 μg/L, 1 μg/L, 9 μg/L, and 27 μg/L treatment groups, respectively. At 24 h (before renewing the exposure solution), mean concentrations of PCP in the exposure solutions were 0.08 μg/L, 0.87 μg/L, 8.47 μg/L, and 32.61 μg/L at 24 h in the 0.1 μg/L, 1 μg/L, 9 μg/L and 27 μg/L exposure groups, respectively. No PCP was detected in the control groups.

DISCUSSION

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. Acknowledgment
  8. REFERENCES

Our results demonstrate that PCP significantly alters plasma thyroid hormone levels, as well as the expression level of selected genes associated with thyroid hormone metabolism and signaling in the HPT axis and liver. The concentrations of PCP we used in the present study (0.1 μg/L, 1 μg/L, and 9 μg/L) have been reported to be prevalent in the aquatic environment [8]; thus, our results suggest that environmental concentrations of PCP might have an adverse effect on fish thyroid systems in an aquatic environment and could affect the development of both adult fish and their offspring.

Although the mortality rate of adult zebrafish was not significantly affected by PCP exposure, a significant decrease in body weight was observed in males exposed to 27 µg/L PCP. These data are in agreement with a previous study, which observed significant effects on growth (total length and body wt) in Japanese medaka (Oryzias latipes) with a 28-d exposure to 200 µg/L PCP [26]. In the present study, the inhibition of growth in males might be partially explained by the significant decrease in plasma T3 levels, as T3 plays an essential role in the regulation of development and growth in fish [27]. In females, gonadal-somatic index was significantly reduced in the highest exposure group (27 µg/L PCP). The low gonadal-somatic index might be related to disruption of the endocrine pathways regulating reproduction, generally as the HPG-axis signaling. Our results indicate that low concentrations of PCP have significant effects on growth of zebrafish and could be endocrine-disrupting chemicals.

We observed an increase in plasma T4 levels after PCP exposure in fish. In contrast, most studies have found reduced T4 levels with treatment at higher doses of PCP in fish [18], rats [13-15], ewes [16], and cell lines [17]. Pentachlorophenol has twice the affinity of T4 to thyroid hormone serum binding transthyretin (TTR) [28], and also affects thyroid hormone metabolism by competitively inhibiting iodothyronine sulfation in vitro [29]. The increases in circulating T4 levels observed in the present study may be attributable to different mechanisms or reflect toxicity of long-term low-dose exposure to PCP. Decreased T3 in males and unaltered T3 in females on exposure to PCP were observed in our study. Although the reason for the sex difference is unclear, a previous study also reported a similar result, in which significant increased TT3 levels and unaltered TT3 levels were observed in PCP-Na–treated female and male rats, respectively [15]. The significant decrease in T3 observed in our study is consistent with previous studies [13, 14], and the observed decrease in T3, at least in part, can be attributed to the thyroid hormone metabolism-disrupting properties of PCP. Indeed, this observation corresponds with the PCP-induced downregulation of dio2 and upregulation of sult1st5 observed in the liver of male fish.

In fish, TSH secretions function as common regulators of the thyroidal axis as feedback mechanisms triggered by changes in the concentration of circulating thyroid hormones [30]. Interestingly, previous studies have reported that changes in tshβ mRNA levels may be related to alterations in T4 levels in fish [31]. In addition, downregulation of tshβ gene transcription has been related to increased levels of T4 in fish after exposure to PBDEs [22]. Consistent with these results, our study revealed that the reduction in tshβ expression could be explained as a negative feedback response to increased levels of T4. Physiological actions of thyroid hormone are usually mediated through interaction with nuclear receptors, and thyroid hormone receptor β (trβ) is one of the main thyroid hormone receptor isoforms. In the present study, trβ gene transcription was significantly downregulated in the male brain and unaltered in female brain. The downregulation of trβ gene expression in the male brain might result from the PCP-induced decline in plasma T3 levels, as the mRNA transcription of trβ was previously reported to be autoinduced in the fish brain by T3 [32].

The TTR is a key thyroid hormone carrier protein that maintains extra thyroidal stores of thyroid hormone, regulates the supply of thyroid hormone to various target tissues, and plays an important role in the thyroid axis in fish [27]. Previous work has demonstrated that phenol compounds such as PCP, halogenated phenols, and tetrabromobisphenol A have strong affinities for TTR [28, 33, 34]. In the present study, ttr gene expression levels were significantly downregulated in the male liver and were upregulated in the female. Our results confirmed that PCP strongly influences TTR expression, as previously reported [28, 33]. Interestingly, the effect of PCP on ttr expression was different in female and male livers, which might be associated with the different T3 levels observed in male and female fish. In addition to TTR, albumin and thyroxine-binding globulin can also bind to thyroid hormones in fish plasma [31]. As such, further studies are needed to better understand the different response to PCP exposure between the sexes.

Hepatic deiodinases are important regulators of circulating and peripheral thyroid hormone levels in vertebrates. In fish, there are 3 dio genes: dio1, dio2, and dio3. Dio1 has a considerable influence on iodine recovery and thyroid hormone degradation [35]. Dio2 is responsible for the conversion of T4 to T3, allowing adequate availability of local and systemic T3 [35], and dio3 is a purely inactivating enzyme [35]. Our results demonstrate that dio1 was significantly downregulated in the liver of both females and males, and the expression of dio2 was upregulated in females but downregulated in males. Therefore, we suggest that the downregulation of dio1 is partially responsible for the increased T4 concentrations we observed. Similarly, Yu et al. [22] demonstrated that exposure to PBDEs downregulated mRNA expression of dio1 and increased plasma T4 concentrations in adult zebrafish [22]. Conversely, the decrease in dio2 expression in males may, at least partly, be associated with the reduced levels of circulating T3, while the upregulation of dio2 expression in females was associated with an increasing trend in T3 levels. The different expression levels of dio2 induced by PCP in male and female fish is consistent with a previous report, which showed that PCP-Na exposure inhibited the dio2 expression of male rats by 79.2% but did not affect the dio2 expression of female rats [15]. In addition, previous studies also provide evidence for sex differences in the magnitude of T3-induced relative mRNA responses for dio3 in liver, as well as dio2 and dio3 in the fish brain [32]. Thus, the cause of the sex differences in dio2 expression may be attributable to the different levels of T3 in females and males. Interestingly, the significant changes in expression of dio1 and dio2 expression may indicate a regulatory role in response to altered thyroid hormone levels and confirm that dio2 is the major contributor to thyroid hormone activation in fish [36].

Uridinediphosphate glucoronosyltransferases and sulfotransferases (SULTs) play important roles in thyroid hormone homeostasis via the major pathway for T4 conjugation [37, 38]. Uridinediphosphate glucoronosyltransferases play a role in decreasing circulating thyroid hormones, and upregulation of ugt gene expression or enzyme activities have generally been observed in rats and zebrafish exposed to different chemicals [23, 39]. In the present study, increased ugt1ab expression could possibly be explained as an autoregulatory response to increased T4 levels, by increased biliary elimination of the conjugated hormone within the thyroid axis. Sulfonation has been viewed as a key step in thyroid hormone metabolism, and it may increase the hydrophilicity and the biliary excretion of the hormone [40]. Sulfonation reactions are catalyzed by SULTs [38]. In zebrafish, sult1 st5 appears to be the only known enzyme that displays substrate specificity exclusively for thyroid hormones and their metabolites [41]. In the present study, we found that sult1 st5 expression was downregulated in female livers and upregulated in male livers, in response to PCP exposure. These results are consistent with a previous study in which the expression of SULT in female and male rare minnow liver were different after PCP treatment [42]. We suggest that downregulation of sult1 st5 in females may be partially responsible for the increased T4 concentrations and unchanged T3 levels, and that the upregulation of sult1 st5 in the male liver might play a role in reducing the circulating levels of T3.

In summary, our results demonstrate that PCP can affect the thyroid endocrine system at lower concentrations in fish. The thyroid hormone disrupting effects of environmental levels of PCP was probably associated with altered thyroid hormone metabolism, as evidenced by marked alterations in the expression levels of dio1, dio2, sult1 st5, and ugt1ab. Interestingly, sex-specific effects on gene expression and plasma TT3 were observed. The causes of these sex-specific differences are unclear and require further study. Nevertheless, our results suggest that analysis of the HPT axis might be suitable for determining thyroid endocrine disruption after PCP exposure.

Acknowledgment

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. Acknowledgment
  8. REFERENCES

This work was supported by grants from the China IWHR Program (HJ1339), the National Natural Science Foundation of China (No. 21247007), and the National Water Program (2012 ZX07203-006).

REFERENCES

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. Acknowledgment
  8. REFERENCES
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