Triclosan,an antimicrobial drug,induced reproductive impairment in the freshwater fish,Anabas testudineus (Bloch,1792)

Introduction

Triclosan (TCS; 5-chloro-2-(2,4-dichlorophenoxy) phenol), a synthetic lipid-soluble antimicrobial agent, is widely used as an antiseptic, disinfectant, or preservative in personal care products, household items, and medical devices (Yueh and Tukey, 2016). In a study, about 15 different pharmaceuticals and personal care products (PPCPs), including triclosan, have been detected in the River Ganges of India, at the nanogram level, suggesting high environmental risks (Singh and Suthar, 2021). The polychlorinated aromatic structure of triclosan permits it to maintain antimicrobial functions, and also allows it to resist degradation, resulting in high persistence in the environment (Bedoux et al., 2012). The chemical stability and widespread use of triclosan with a half-life of less than 1 h in abiotic conditions, 8 h in the atmosphere, up to 10 days in lake water, and 18 days in clay loam soil, paves the way to enter into the aquatic environment (Kookana et al., 2011; Ying et al., 2007). In the aquatic ecosystem, triclosan has been shown to accumulate largely in sediments due to low water solubility thereby causing toxicity to aquatic organisms (Grove et al., 2003; Jug et al., 2011). Besides the antimicrobial property, triclosan also possesses endocrine-disrupting effects that have been established in various fish species (Raut and Angus, 2010; Wang et al., 2017, 2018, 2022).

Estrogenic effects of similar compounds are known to affect the reproductive, developmental, and endocrine functions in vertebrates. Estrogens are the key regulators of many physiological functions associated with reproduction including reproductive behavior, steroidogenesis, endocrine functions, and sexual differentiation. The gene for aromatase, an enzyme responsible for the biosynthesis of estrogen, is expressed in the brain along with estrogen, which is involved in the regulation of neurogenesis and sexual behavior (Cheshenko et al., 2008; Diotel et al., 2010; Lephart, 1996). In teleosts, the two isoforms of the aromatase gene, namely, cyp19a1a or aromatase A and cyp19a1b or aromatase B are predominately expressed in gonads and radial glial cells of the brain, respectively (Diotel et al., 2010; Piferrer and Blázquez, 2005). Aromatization is also responsible for brain sexualization in fish possibly by stimulating pituitary gonadotropins, which in turn regulate the level of estrogen by the up-regulation of cyp19a1a gene expression (Senthilkumaran et al., 2015; Zubizarreta et al., 2020). Expression of the aromatase A gene is also involved in sexual differentiation and the formation of inter-sex in the fish by catalyzing aromatization, the transition of androgen to estrogen (Guiguen et al., 2010; Urbatzka et al., 2012). Thus the biosynthesis of estrogens in the brain undergoes a positive auto-regulatory feedback loop leading to a dramatic increase in aromatase expression in sexually mature fish, with elevated levels of circulating steroids (Coumailleau et al., 2015).

In teleosts, reproduction is controlled by gonadal steroids, namely, estrogen and androgen along with the pituitary gonadotropins as luteinizing hormone and follicle-stimulating hormone, where the ratio of the gonadal steroids determines sexual differentiation and maintenance of the reproductive physiology in fish (Zhang et al., 2013). Steroidogenesis is initiated by steroidogenic acute regulatory protein (StAR), which mediates the rate-limiting transport of cholesterol into the mitochondria where it gets converted to pregnenolone (Arukwe, 2008; Stocco, 2000). The key steroidogenic enzymes, namely, 3β-hydroxysteroid dehydrogenase (3β-HSD) and 17β-hydroxysteroid dehydrogenase (17β-HSD) are involved in gonadal steroidogenesis by active biosynthesis of progesterone and testosterone, regulated by gonadotropins (Ings and Van Der Kraak, 2006; Nakamura et al., 2009; Rasmussen et al., 2013). The catalytic action of several steroidogenic enzymes accounts for the production of androgens, which finally get converted into estrogens by the catalytic activity of the aromatase enzyme.

Aromatase gene expression and histological analysis are together considered valuable biomarkers for evaluating the reproductive toxic effects of the chemicals. Some studies have suggested that the endocrine-disrupting effects of triclosan occur by interfering with the hypothalamic-pituitary-gonadal axis (Dann and Hontela, 2011; Song et al., 2020; Wang et al., 2018). Some of the mammalian in vitro studies have reported that triclosan exerted estrogenic effects by binding with estrogen receptors (Gee et al., 2008; Huang et al., 2014; Wang and Tian, 2015). Despite this, a recent study also examined the ability of triclosan to interfere with estrogen receptor transcriptional activity using zebrafish-specific in vitro and in vivo reporter gene assays (Serra et al., 2018). Chemically, triclosan is chlorinated biphenyl ethyl, which shares structural similarities to other persistent endocrine-disrupting chemicals such as polychlorinated biphenyls, bisphenol A, and dioxins, and also with the endogenous thyroid hormone (Crofton et al., 2007). Chronic exposure to PPCPs has been shown to disturb the hypothalamic-pituitary-gonadal axis in zebrafish (Hamid et al., 2022). In this context, the present study was designed to assess the reproductive toxicity of triclosan in the freshwater fish Anabas testudineus using measurable endpoints such as steroidogenic enzyme activities, aromatase gene expression, levels of serum hormones, and histology of gonads.

Materials and methods

Design of experiments and sample processing

During the pre-spawning period, there was no remarkable sexual dimorphism thus for grouping the test animal, sex distinction was made only based on the large belly size of females, and others were grouped as males. Accordingly, fish maintained in the stock were grouped in separate glass tanks consisting of 15 fish per group per sex in duplicates, expecting 10 fish of the same sex in each group on dissecting the animal at the end of the exposure period. Triclosan at three different concentrations was exposed to the fish for 90 days along with control groups, which include the toxicant-free (only tap water) and solvent-control (DMSO; 0.001% v/v) group. The selection of test concentrations was based on previous literature, in which two environmentally relevant concentrations (0.009 and 9 μg L⁻¹) representing the lowest doses detected in the surface waters of Indian rivers (Nag et al., 2018; Ramaswamy et al., 2011), and one-tenth of the LC₅₀-96 h, that is, 176.7 μg L⁻¹ representing the sublethal concentration (Priyatha and Chitra, 2018).

The experiments were designed, grouped, and performed in a static renewal system holding 40 L of water in a glass tank (Table 1). The water was renewed every 96 h for both control and treatment groups by siphoning 80–90% of the water and refilling with the same volume of water and test concentrations in the respective tanks. The concentrations of triclosan were selected as per the previously published data (Nag et al., 2018; Priyatha and Chitra, 2018; Ramaswamy et al., 2011). The particulate wastes deposited at the bottom of the tank were removed carefully without giving stress and disturbance to the fish. All biochemical data represented in the figures denote the mean values of 10 animals per group, in duplicates.

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Table 1.

The experimental design and grouping used in the study.

Group	Fish in the pre-spawning stage (April–June) n = 10/group, in triplicate, maintained for 90 days
Control	No toxicant (only tap water)
Vehicle	Solvent - dimethylsulfoxide (DMSO; 0.001 % v/v)
Treatment (Triclosan)	0.009 μg L−1
	9 μg L−1
	176.7 μg L−1

Biochemical analysis

Gonadal steroidogenic enzymes

The activities of 3β-hydroxysteroid dehydrogenase (3β-HSD) and 17β-hydroxysteroid dehydrogenase (17β-HSD) were measured in the gonads according to Bergmeyer (1974); with minor modifications (Bergmeyer, 2012). Briefly, the gonadal homogenate (10%) was prepared in 5 mM potassium phosphate buffer containing 1 mM EDTA and 20% spectroscopic grade glycerol, and the supernatant was collected after centrifugation at 10,000 x g for 10 min at 4°C. For the assay of 3β-HSD, an aliquot of the supernatant was mixed with pyrophosphate buffer (100 mM), NAD (0.5 mM), and dehydroisoandrosterone (0.1 mM), and absorbance was read immediately at 340 nm for 5 min at 30 s interval in a spectrophotometer against the blank. The units are expressed in μmol of NAD reduced/min/mg protein. The assay mixture of 17β-HSD containing pyrophosphate buffer (100 mM), NADPH (0.5 mM), and 1,4-androstenedine-3,17-dione (0.8 mM) was read at 340 nm immediately after the addition of sample at 30 s interval for 5 min in a spectrophotometer against the blank. The enzyme activity was expressed as μmol of NADP formed/min/mg protein.

Aromatase enzyme in the brain and gonads

The concentration of fish aromatase enzyme in the brain and gonads was assayed using fish aromatase ELISA kits, following the manufacturer’s instructions. The aromatase standards and samples were added to the wells of ELISA plates pre-coated with fish aromatase antibodies. Followed by the addition of biotin conjugate and enzyme conjugate reagents, the plate was incubated at 37°C for 60 min. After incubation, the plates were rinsed 5 times with wash solution, then chromogenic reagent was added and the absorbance was measured at 450 nm within 15 min. Total protein was estimated using bovine serum albumin as the standard (Lowry et al., 1951).

Histology of gonads

The control and triclosan-exposed gonads, ovary, and testis collected after 90 days were fixed in 10% buffered formalin for 24–48 h, which were later used for histological analysis. The tissues were dehydrated in ascending grades of alcohol and cleared in xylene until they became translucent. Then the gonadal tissues were embedded in molten paraffin wax for an hour for the complete impregnation to make the tissue blocks. Tissue sections of 4 to 6 μm thickness were made using a rotary microtome, double-stained with hematoxylin and eosin, and finally mounted using DPX (Roberts, 2012). Five slides of gonadal tissues were prepared from each group and examined under a light microscope equipped with a digital camera (Magcam DC series camera). Histological fields showing lesions were focused by a random selection of frames that varied between 5 and 6 frames and photographed at 40–100× magnification for the ovary, and 100–400× magnification for the testes. Histological lesions of the ovary and testis were confirmed as per the criteria previously described in the OECD guidelines (Johnson et al., 2009).

Aromatase (cyp19a1a and cyp19a1b) gene expression

Total RNA was extracted manually from each sample of approximately 50 mg tissue using TRIzol reagent. Briefly, tissue was homogenized and incubated with TRIzol reagent at room temperature for 10 min. Chloroform was added to the mixture and centrifuged at 13525 x g for 15 min at 4°C. The supernatant collected was incubated with ice-cold isopropanol for 10 min at 4°C followed by centrifugation at 13525 x g for 10 min at 4°C. The pellets were collected in 75% ethanol, air-dried, and dissolved in RNAase-free water (30–50 μL), later stored at −20°C for further procedures. The quality and concentration of total RNA extracted were estimated by the NanoDrop one Microvolume UV-Visible Spectrophotometer (Thermo Fisher Scientific).

The first-strand cDNA was synthesized from the isolated RNA using Thermo Scientific RevertAid First-Strand cDNA Synthesis Kit according to the manufacturer’s instructions. RNA (0.1 ng – 5 µg) was reverse transcribed to cDNA in a final reaction volume of 20 µL. The samples were then incubated at 42°C for 60 min, and the reaction was terminated by heating at 70°C for 5 min. The obtained cDNA samples were used for the qPCR analysis using the housekeeping gene gapdh as the internal control. The selection of specific primers for the amplification of aromatase (cyp19a1a and cyp19a1b) and gapdh genes were based on the previous report as shown in Table 2 (Pal et al., 2018).

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Table 2.

List of primers for cyp19a1a and cyp19a1b transcripts.

Gene		Primer sequence (5′-3′)
cyp19a1a	Forward	TGACAGATGCTCAGGGACAG
	Reverse	CTGTGAAGCCAGTCCAGTCT
cyp19a1b	Forward	TGCTCTGGAGGAAAGTGAGG
	Reverse	TCAGCAGATTGAGAGCGTCC
gapdh	Forward	AGGGGCTCAGTATGTTGTGG
	Reverse	AGGAGGCATTGCTGACAACT

qPCR was performed using the SYBR Green PCR master mix (Takara Bio INC), and the reaction was carried out by using the QuantStudio 3 Real-Time PCR System (Applied Biosystems, USA). For each qPCR reaction, 1 μL of cDNA (∼50 ng) was mixed with 12.5 μL 2X SYBR Green PCR master mix (Takara Bio INC) and 5 pmol each of forward and reverse primers in a final volume of 25 μL. The reactions were amplified for 2 min at 50°C, 10 min at 95°C, 0.30 min at 95°C, and 0.45 min at 50°C for 40 cycles. Amplification was followed by thermal denaturation for 0.15 min at 95°C, 0.15 min at 50°C, and 0.15 min at 95°C, and the melting curves generated were used to verify the amplification specificity. The housekeeping gene gapdh was used as the reference gene. The gene expression of aromatase relative to the gapdh gene was estimated based on the comparative C_T method (ΔΔC_T). The relative gene expression was measured based on the equation log2RQ, where RQ was calculated as 2^−ΔΔC_T, and ΔΔC_T = (C_{T, target RNA} – C_{T, reference RNA}) – (C_{T, calibrator} – C_{T, reference RNA}) (Livak and Schmittgen, 2001; Winer et al., 1999).

Results

Body weight, mucus deposition, and organ weights

The body weight of both male and female fish after 90 days of triclosan exposure at 0.009, 9, and 176.7 μg L⁻¹ concentrations showed a significant (p < .05) reduction in a concentration-dependent manner when compared with the corresponding control groups (Figure 1(A)). However, the percentage of mucus deposition in the body of fish increased significantly (p < .05) with an increase in the concentration of triclosan exposure (Figure 1(B)). The absolute and relative weights of gonads, and the weight of brain tissue showed significant (p < .05) reduction in all treatment groups of both sexes (Figure 1(C)–(E)).

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Figure 1.

Effect of triclosan on (A) the body weight; (B) Mucus deposition; (C) Weights of gonads; (D) Gonadosomatic index (GSI); and (E) Weight of brain tissues in the fish, Anabas testudineus (Mean ± SD; n = 10/group, in triplicates; Asterisks (*) denotes significant differences against the control groups).

Gonadal steroidogenic enzymes

The activities of gonadal steroidogenic enzymes, namely, 3β-hydroxysteroid dehydrogenase (3β-HSD) and 17β-hydroxysteroid dehydrogenase (17β-HSD) were decreased significantly (p < .05) in the ovary and testis after 90 days of all tested concentrations, which was concentration-dependent on comparison with the respective control groups (Figure 2(A)).

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Figure 2.

Effect of triclosan on the (A) Activities of steroidogenic enzymes (B) Concentration of aromatase enzyme; (C) Aromatase Cyp19a1a gene expression; (D) Aromatase Cyp19a1b gene expression in the fish, Anabas testudineus (Mean ± SD; n = 10/group, in triplicates; Asterisks (*) denotes significant differences against the control groups).

Serum hormones

The levels of serum follicle-stimulating hormone (FSH) and luteinizing hormone (LH) showed a significant (p < .05) reduction in both male and female fish in a concentration-dependent manner than the corresponding control groups (Figure 3(a)). The level of estradiol in the serum of female fish declined significantly (p < .05) whereas increased significantly (p < .05) in male fish (Figure 3(b)). However, a significant (p < .05) reduction in the level of testosterone was observed after triclosan exposure in the serum of both male and female fish when compared to the respective control groups (Figure 3(b)).

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Figure 3.

Effect of triclosan on the levels of (A) Gonadotropic hormones; (B) Sex steroid hormones in the serum of the fish, Anabas testudineus (Mean ± SD; n = 10/group, in triplicates; Asterisks (*) denotes significant differences against the control groups).

Histology of gonads

Exposure of triclosan at 0.009, 9 and 176.7 μg L⁻¹ concentrations for 90 days showed remarkable histological lesions in the pre-spawning gonads of A. testudineus when compared with the control tissues. Histomorphology of ovarian tissue in the control groups (negative-control and vehicle-control) showed vitellogenic oocytes within well-developed, compact ovarian lamellae (Figures 4(A) and 4(B)). Fish exposed to 0.009 μg L⁻¹ concentration of triclosan showed loosely arranged immature oocytes (Figure 4(C)). In the 9 μg L⁻¹ concentration group, the membrane of oocytes showed prominent blebbing as a sign of oocyte regression, and the oocytes were immature and loosely arranged (Figure 4(D)). In the sublethal exposure group, that is, at 176.7 μg L⁻¹ concentration of triclosan, there were no vitellogenic oocytes but the follicles were formed of the developing chromatin nucleolus and perinucleolus stages along with the aggregation of melanomacrophage centers (Figure 4(E)). The major pathologies observed in the pre-spawning oocytes after 90 days of triclosan exposure includes loosely arranged oocytes, empty follicles, RBC infiltration, and melanomacrophage aggregation (Figure 5(A)–(D); Table 3).

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Figure 4.

Photomicrographs showing pre-spawning ovary (A-E) and testis (a-e) of the fish Anabas testudineus. A-Control; B-Vehicle showing vitellogenic oocytes; C-Triclosan at 0.009 μg L⁻¹ for 90 days, loosely arranged immature oocytes (←); D-Triclosan at 9 μg L⁻¹ for 90 days, loosely arranged immature oocytes (*), membrane blebbing (←); E-Triclosan at 176.7 μg L⁻¹ for 90 days, melanomacrophage aggregation (←); a-Control; b-Vehicle showing compact seminiferous tubules; c, d, e-Triclosan-exposed groups (0.009, 9 and 176.7 μg L⁻¹ for 90 days, respectively), Vacuolization (V), Loss of spermatozoa (*).

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Figure 5.

Photomicrographs showing representative pathologies of pre-spawning ovary (A-D) and testis (a-d) in Anabas testudineus. A-Loosely arranged oocytes; B-Empty follicles; C-RBC infiltration (*); D-Melanomacrophage aggregation (*); a-Loss of spermatozoa; b-Vacuolization; c-Dedifferentiated seminiferous tubules; d-Melanomacrophage aggregation (←).

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Table 3.

Semi-quantitative scoring of pathological changes in the ovary of Anabas testudineus after triclosan exposure.

Histological lesions	Triclosan
	0.009 μg L−1	9 μg L−1	176.7 μg L−1
Loosely arranged oocytes	+	++	+++
Membrane blebbing	+	++	+++
Aggregation of melanomacrophage centers	-	-	+++
Empty follicles	-	-	+++
RBC infiltration	-	-	+++

- Nil, + Mild, ++ Moderate, +++ Severe.

The histological examinations in the testis of the fish after triclosan exposure also revealed some severe abnormalities. The control testis of the pre-spawning phase was normal showing compact seminiferous tubules with orderly arranged different spermatogenic stages (Figures 4(A) and (B)). The changes observed in testicular morphology after 0.009 μg L⁻¹ concentration of triclosan were the development of vacuolization in the interstitial cells and within the seminiferous tubules (Figure 4(C)). On exposure to 9 μg L⁻¹ concentration of triclosan for 90 days developed thickening in the wall of seminiferous tubules with vacuolization and loss of spermatozoa (Figure 4(D)). The regression of seminiferous tubules, thickening of the interstitium with massive destruction in germ cell epithelium, and loss of spermatozoa were identified after 176.7 μg L⁻¹ concentration of triclosan exposure (Figure 4(E)). Some of the significant morphological aberrations in the testis included loss of spermatozoa, vacuolization, dedifferentiated seminiferous tubules by the formation of the thickened interstitium, regression of the tubules, and development of melanomacrophage aggregation (Figure 5(A)–(D); Table 4).

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Table 4.

Semi-quantitative scoring of pathological changes in the testis of Anabas testudineus after triclosan exposure.

Histological lesions	Triclosan
	0.009 μg L−1	9 μg L−1	176.7 μg L−1
Vacuolization	+	++	+++
Thickening in the wall of seminiferous tubules	-	++	+++
Loss of spermatozoa	-	++	+++
Regression of seminiferous tubules	-	-	+++
Thickening of the interstitium	-	-	+++
Destruction of testicular epithelium	-	-	+++
Melanomacrophage aggregation	+	++	-

- Nil, + Mild, ++ Moderate, +++ Severe.

Discussion

Some environmental contaminants possessing estrogenic properties like pesticides, pharmaceuticals, personal care products, plasticizers, and synthetic estrogens enter the aquatic environment through untreated wastewater disposal, rainwater run-off, transportation, accidental spillage, and groundwater discharge thereby affecting the life of aquatic organisms, including fish (Holt, 2000; Müller et al., 2020; Shincy and Chitra, 2020). Triclosan, an emerging contaminant, having antimicrobial properties has been detected in soils and sediments at 2.7 to 26.8 μg L⁻¹ (McAvoy et al., 2002; Reiss et al., 2002; Waltman et al., 2006), in groundwater and surface water at 0.055–0.184 and 0.041–0.077 μg L⁻¹, respectively (Das Sarkar et al., 2020), and also in aquatic life and humans (Chalew and Halden, 2009; Weatherly and Gosse, 2017). In the Indian scenario, the presence of triclosan has been reported in the surface water at the mean concentration of 944 ng L⁻¹ in the Tamiraparani River, Tamil Nadu, and with the highest concentrations of 5160 ng L⁻¹ in the Cheranmahadevi followed by Tirunelveli (3800 ng L⁻¹), Kaveri (40.7 ng L⁻¹) and Vellar (8.95 ng L⁻¹) rivers, respectively (Ramaswamy et al., 2011). In the sediments, triclosan concentrations ranged between 132 and 3073 ng kg⁻¹ from the samples collected from the Valliyar estuary, Kanyakumari district, Tamil Nadu (Jeyakumar et al., 2012), and in water (1.1–9.65 μg L⁻¹), sediments (5.11–50.36 μg kg⁻¹) and fish samples (13 to 1040 μg kg⁻¹) from the River Gomti, a major tributary of the River Ganga (Nag et al., 2018). The quantification of triclosan in the freshwater fish, Gibelion catla from the Kaveri River, India was found in the range between 0.73 and 50 ng/g wet weight (Shanmugam et al., 2014), which revealed the persistence of the compound throughout the Indian rivers. The ubiquitous presence of triclosan in aquatic bodies and its endocrine-disrupting potential stand significant to assess its role in the reproduction of the fish, Anabas testudineus.

In the current study, triclosan exposure to the fish for 90 days during the pre-spawning period at three different concentrations 0.009, 9, and 176.7 μg L⁻¹ caused changes in many aspects of fish physiology including the growth, weights of gonads, and brain tissues, and other reproductive and hormonal parameters like steroidogenic enzyme activities, aromatase gene expression, serum hormone levels, and histology of gonads. Alterations in the body weight and organ weights are the common parameters used to generalize the direct toxic effects of the compound. Triclosan exposure reduced the body weight, brain tissue weight, and absolute and relative weights of gonads in a concentration-dependent manner on both sexes when compared with the respective control groups thereby illustrating the systemic effects of the toxicant. In most ecotoxicological studies, the evaluation of body weights and tissue weights are considered one of the most sensitive indicators of toxicant exposure where these changes are associated with treatment-related anorexia and/or cellular and tissue damages (Bailey et al., 2004; Sumi and Chitra, 2020). The gonadosomatic index or the relative weight of the gonads is considered a good indicator to measure reproductive activity upon exposure to estrogenic compounds (Sumi and Chitra, 2020). Triclosan exposure significantly affected the gonadosomatic index in both sexes consistent with a previous study after sublethal exposure of chlordecone for 30 days to the cichlid fish, Pseudetroplus maculatus (Asifa and Chitra, 2019). Nile tilapia Oreochromis niloticus exposed to nonylphenol, an estrogenic compound, has been shown to cause a decline in GSI, which was found associated with histological lesions in the ovary, and reduction in the levels of estradiol and vitellogenin in the female fish (El-Sayed Ali et al., 2014). Similarly, male Nile tilapia exposed to diuron and its metabolites saw a reduction in GSI and altered testicular histology with a profound reduction in the levels of testosterone and 11-ketotestosterone thereby suggesting anti-androgenic activity of the herbicide (Pereira et al., 2015). In the current observations, reduction in female and male GSI suggests that triclosan possesses both estrogenic and anti-androgenic properties.

The skin of teleost fish is unique as it secretes mucous that is involved in immune functions (Salinas et al., 2011), and also plays a vital role in the first-line of defense against the invasion of pathogens or the entry of environmental contaminants (Groff, 2001). Chronic exposure to triclosan elevated the percentage of mucous deposition as the onset of the primary defensive mechanism to prevent or escape from the toxicant. A similar stress-induced mucous deposition has been observed in the fish Oreochromis niloticus after triclosan exposure. (Vijitha et al., 2017) Anabas testudineus exhibited an incidental defensive mechanism on exposure to the environmental contaminant fullerene C₆₀ nanomaterial for 60 days (Sumi and Chitra, 2019).

In the steroidogenic pathway, the major oxidoreductases, namely, 3β-hydroxysteroid dehydrogenase (3β-HSD) and 17β-hydroxysteroid dehydrogenase (17β-HSD) enzymes are involved in the synthesis of bioactive steroids (Sampath Kumar et al., 2000). 3β-HSD converts Δ5 into Δ4 steroids where pregnenolone and dehydroepiandrosterone (DHEA) are converted to progesterone and androstenedione, respectively, whereas 17β-HSD reduces 17-ketosteroids into 17-hydroxysteroids so that androstenedione is converted into testosterone (Mindnich et al., 2004). Hence both enzymes actively participate in the biosynthesis of sex steroids, and they are widely used as biomarkers to assess gonadal steroidogenesis in fish. In the present study, triclosan exposure caused a reduction in the activities of 3β-HSD and 17β-HSD enzymes in the ovary and testis of the fish in a concentration-dependent manner. The findings suggested that triclosan disrupted gonadal steroidogenesis in the fish, which ultimately lead to altered biosynthesis of gonadal hormones. The present results were in agreement with another study that sublethal exposure of dibutyl phthalate to the freshwater fish, Pseudetroplus maculatus, decreased the activities of 3β-HSD and 17β-HSD in the ovarian and testicular tissues thereby suggesting a disruption in the gonadal steroidogenic pathway (Sajla et al., 2019; Sruthi et al., 2021).

Aromatase, the enzyme that converts testosterone into estradiol is associated with the maintenance of reproductive behavior (Cheshenko et al., 2008; Huffman et al., 2013), steroidogenesis (Tomy et al., 2007), and sexual differentiation in fish (Kwon et al., 2001). The present study observed that triclosan exposure elevated the concentration of aromatase enzyme in the brain and testis of A. testudineus indicating the auto-regulatory feedback loop driven by estradiol. However, the enzyme concentration was decreased in the ovary, and this could be due to the difference in the sensitivity of ovarian tissue toward the estrogenic compound. The concomitant decline in the level of estrogen in the female fish indicated failure of aromatization in the ovarian tissue. Further, the study observed that brain aromatase level was higher than in the gonadal tissues. The current findings demonstrated that triclosan equally targeted brain and gonadal aromatase activity thereby revealing the endocrine-disrupting effects of the compound. Similar potential targets on the activities of the brain and gonadal aromatase enzyme have been documented after exposure to endocrine-disrupting chemicals in zebrafish, Danio rerio (Hinfray et al., 2006). Likewise, inhibition in the aromatase concentration has been observed in the ovary of the marine fish cunner Tautogolabrus adspersus after exposure to endocrine-disrupting chemicals such as estradiol and ethynylestradiol (Mills et al., 2014).

In teleosts, the aromatase genes, namely, cyp19a1a or aromatase A, and cyp19a1b or aromatase B are structurally and functionally different isoforms expressed in the ovary and brain, respectively (Piferrer and Blázquez, 2005). Aromatase or estrogen synthetase is a member of the cytochrome P450 family involved in the biosynthesis of estrogen and contributes to gonadal sex differentiation (Guiguen et al., 2010). The current study observed that triclosan exposure up-regulated cyp19a1a and cyp19a1b aromatase gene expression in the brain of both sex and testis, while down-regulated in the ovary thereby revealing a tissue-specific differential gene expression. The up-regulated gene expression of cyp19a1a and cyp19a1b in the brain and testis of male fish suggested an increased estrogen synthesis during the pre-spawning period as evidenced by the rise in the level of serum estradiol. Estrogen is also involved in the neural aromatase expression through estrogen-responsive elements (EREs) in the promoter region of the cyp19a1b gene (Ramachandran et al., 1999). Besides, the presence of several putative cis-regulatory elements including an androgen-responsive element (ARE) found in the proximal promoter sequence of cyp19a1b genes makes the promoter sensitive to androgens as well (Diotel et al., 2010; Sawyer et al., 2006). However, the promoter region of cyp19a1a contains a steroidogenic factor-I (SF-1) regulatory element, which is less responsive to estrogen but activated by gonadotropins to increase cyp19a1a expression for estradiol production (Das and Mukherjee, 2013). The results were consistent with a previous study that showed an up-regulation of aromatase gene expression in the gonads and hypothalamus of juvenile male Yellow River carp, Cyprinus carpio exposed to triclosan, which has been mediated through the anti-androgenic mode of action (Wang et al., 2018).

In female fish, triclosan exposure caused up-regulation in gene expression of both aromatases in the brain while down-regulated in the ovary. The high affinity of triclosan towards the receptor of cyp19a1a gene, and chronic exposure could have resulted in receptor saturation thereby down-regulated the gene expression in the ovary (Li et al., 2017). A similar observation has been found in the ovary and brain tissues of zebrafish after bisphenol A exposure (Risalde et al., 2021). Aromatase expression is a good biomarker to test estrogen level as well as reproductive behavior (Guyón et al., 2012) where triclosan-induced modulation in the gene expression could affect neurodevelopment resulting in behavioral modifications (Schultz et al., 2012). Thus the variability in aromatase expression observed in the current study serves as an indicator of reproductive impairment in the fish, A. testudineus.

It is well-known that any temporal changes in the steroidogenic enzyme activities could alter the levels of sex hormones and the circulating gonadotropins. As expected, the present findings showed that triclosan exposure in female fish during the pre-spawning period of reproduction caused a concentration-dependent decrease in the levels of serum FSH, LH, estradiol, and testosterone. The results suggested that triclosan regulated sex hormone production by inhibiting the activity of ovarian steroidogenic enzymes. Besides, a low concentration of aromatase enzyme in the ovary could have prevented the conversion of androgen to estrogen thereby reducing the levels of sex hormones in female fish. The decline in the circulating levels of gonadotropins such as FSH and LH after triclosan exposure mediated through the negative feedback on the hypothalamic-pituitary-gonadal (HPG) axis also contributes to the decline of sex hormones (Redding and Patiño, 1993). Zebrafish exposed to triclosan up-regulated gonadotropin and steroidogenic genes perturbed the HPG axis pathway in females, whereas down-regulated expressions of vtg, cyp17, and 17βhsd genes in male fish (Hamid et al., 2022).

Male fish exposed to triclosan in the pre-spawning period caused a decrease in the levels of serum FSH, LH, and testosterone while the level of estradiol increased in a concentration-dependent manner. The inhibition in the testicular steroidogenic enzymes after triclosan exposure had down-regulated the release of gonadotropins thereby inhibiting testosterone secretion by blocking the negative feedback either at the hypothalamus or pituitary, which indicated anti-androgenic effects of triclosan. Meanwhile, triclosan exposure increased the estrogen level in male fish and this could be due to the increase in aromatase enzyme activity by interfering with the male reproductive axis, which indicated the potential non-estrogen receptor-mediated anti-androgenic effects of the toxicant. A similar mechanism of hormonal imbalance has been reported in the juvenile male Yellow River carp Cyprinus carpio confirming the anti-androgenic effects of triclosan in male fish (Wang et al., 2018). The auto-regulatory loop involving estradiol production from testosterone by enhancing the activity of brain aromatase may lead to low sperm counts and decreased sperm quality, which is suspected to affect male fertility in fish (Tang et al., 2017).

Histopathological examinations in the ovary of the pre-spawning phase of the fish revealed severe lesions including loosely arranged immature oocytes, formation of membrane blebbing as a sign of oocyte regression, and absence of vitellogenic oocytes. The ovarian follicles were predominantly observed with the developing chromatin nucleolus and perinucleolus stages, while some follicles appeared empty with RBC infiltration and aggregation of melanomacrophage aggregation. Research has illustrated that gonadal histopathology has gained increasing interest as a diagnostic endpoint for the identification of morphological alterations, and interaction of various physiological and endocrine functions. There is a close relationship between the gondosomatic index and histology of the ovary, as the observed immature and loosely arranged regressed oocytes could be the reason for the reduction of GSI in the female fish. Similar histomorphological abnormalities have been reported in the developing oocytes of zebrafish exposed to estrogenic chemicals (Weber et al., 2003). The characteristic changes observed in the oocytes after triclosan exposure suggests that the ovarian tissues shifted towards various stages of degeneration in the pre-spawning phase that could eventually lead to failure of spawning and reproductive functions.

The tissue clusters of phagocytes function as reservoirs for breakdown products from dead or damaged cells and the site of antigen presentation is called melanomacrophage aggregation (MMAs) (Agius and Roberts, 2003). They are usually found in hemopoietic tissues like the liver and spleen and also in kidneys. In gonads, MMAs are associated with gonadal regression, that is, degradation and resorption of unspawned oocytes and sperm, respectively (Ravaglia and Maggese, 1995). In the present study, MMAs were more prominent in the 176.7 μg L⁻¹ triclosan-treated fishes showing regressed ovaries. While, in males, with the highest concentration of triclosan at 176.7 μg L⁻¹ the testicular maturity was greatly reduced and did not elicit MM aggregation. But at lower concentrations, they were found in abundance. In a previous study in red mullet, Mullus barbatus, exposed to polluted water showed MMAs in ripe testes only, whereas they occurred only in regressing ovaries (Micale et al., 2019). These findings suggest that the presence of ovarian MMAs in Aanabas testudineus is most likely related to ovarian regression and the absence of testicular MMAs suggests reduced testicular maturity and testicular regression following triclosan toxicity. Further, altered water quality (Patiño et al., 2003) and decreased immune response due to the immunosuppressive mechanism of triclosan (Bera et al., 2020) may contribute to the absence of testicular MMAs at higher concentrations.

Triclosan exposure also caused remarkable histomorphological changes in the testis including vacuolization, thickening in the wall of seminiferous tubules, loss of spermatozoa, thickened interstitium, regression of the tubules, and formation of melanomacrophage aggregation. The severity of morphological alterations of the testicular tissues after triclosan exposure was found to progress based on an increase in the concentration of the toxicant. The results were in agreement with another study that showed a concentration-dependent increase in the severity of testicular lesions in zebrafish after exposure to nonylphenol and ethinylestradiol (Weber et al., 2003). Thus, the negative effects in male gonads also contributed to the reproductive toxicity of triclosan in the fish, Anabas testudineus. Besides all, Triclosan exposure may occur through dermal or inhalation routes at workplaces where it is produced as well as in the healthcare industry where it is repeatedly used (Jones et al., 2000; MacIsaac et al., 2014). In a previous study, the effect of triclosan on estradiol production and human CYP19A1 activity in JEG-3 cells was found at an IC₅₀ value of 6.26 μM triclosan showing reduced estradiol production and competitively inhibited CYP19A1 by binding to the steroid-binding pocket (Li et al., 2017). Thus, the handling of triclosan should be done with utmost care as it could elicit similar adversities revealed in our study after prolonged exposure in humans.

Conclusions

The current research provided an overview of the reproductive toxicity of triclosan in the pre-spawning gonads of the freshwater fish, Anabas testudineus. Briefly, this study provided direct evidence that triclosan possesses both estrogenic and anti-androgenic properties to cause reproductive dysfunction in the fish A. testudineus. However, further multigenerational studies are needed for a better understanding of the possible mechanism of action, and sensitivity of gonadal tissues, as the primary target of triclosan toxicity.