Skip to main content
NCBI home page
NIHPA Author Manuscripts logoLink to NIHPA Author Manuscripts
. Author manuscript; available in PMC: 2023 Jan 1.
Published in final edited form as: Neurotoxicol Teratol. 2021 Dec 9;89:107055. doi: 10.1016/j.ntt.2021.107055

Transient developmental exposure to tributyltin reduces optomotor responses in larval zebrafish (Danio rerio)

Rachel C Bernardo 1,2, Victoria P Connaughton 1,*
PMCID: PMC8755603  NIHMSID: NIHMS1763854  PMID: 34896240

Abstract

This study determined the effects of transient developmental exposure to tributyltin (TBT), a well-known anti-estrogenic environmental endocrine disrupting compound, on visual system development of larval zebrafish (Danio rerio). Zebrafish were exposed to either 0.2 μg/L or 20 μg/L TBT for 24 hours when they were aged 24 hours postfertilization (hpf), 72 hpf, or 7 days (d)pf. Immediately after exposure, larvae were transferred to system water for seven days of recovery followed by behavioral testing (startle and optomotor responses) and morphological assessment. TBT-treated larvae displayed age-dependent changes in morphology characterized by delayed/reduced growth and susceptibility to exposure. TBT exposure reduced the number of larvae displaying optomotor responses regardless of age of exposure; eye diameter was also decreased when exposure occurred at 24 hpf or 7 dpf. Startle responses were reduced only in TBT-treated larvae exposed when they were 24 hpf, suggesting transient TBT exposure during the early larval period may cause vision-specific effects.

Keywords: endocrine disruptor, aromatase, OMR, vision, estrogen

1. Introduction

Visual system development is a complex process involving the coordination of cues from both external (environment) and internal (cell signaling) sources. In zebrafish, this process begins at 12 hours post fertilization (hpf) with evagination of optic primordia (Schmitt and Dowling, 1999). By 24 hpf the eyecup is well formed and, at 36 hpf, embedded in the skull (Schmitt and Dowling, 1994). Internally, the first ganglion cells form at ~32 hpf, followed by photoreceptors (50–55 hpf) and other retinal neurons, so that at hatching (72 hpf) all retinal cell types are present (Schmitt and Dowling, 1999). Initial molecular analysis identified an array of genes involved in retinal development, eye formation (Macdonald and Wilson, 1996), and axonal pathfinding (Baier et al., 1996; Karlstrom et al., 1996; Trowe et al., 1996). More recent studies have identified additional cell specific genes, adhesion molecules, and intracellular pathways. Given this complexity, alteration(s) to any aspect of eye/retinal development can cause detrimental effects.

Endocrine disrupting compounds (EDCs) are a class of chemicals that have been linked to obesity (Hatch et al., 2010), metabolic syndrome (Casals-Casas and Desvergne, 2011), Type II diabetes (Alonso-Magdalena et al., 2011), and neurological disorders (Kajta and Wojtowicz, 2013). Endocrine disruptors are diverse in structure (Diamanti-Kandarakis et al., 2009) and ubiquitous in our environment (Schug et al., 2015). The action of most identified EDCs target estrogen pathways (TEDX, 2011) and, consequently, have potent effects on reproductive physiology. In addition, estrogenic and anti-estrogenic EDCs are critical modulators of neural development, including pathways of the visual system (de Paulo et al., 2020; Dong et al., 2006; Hamad et al., 2007; Wang and Huang, 1999). These compounds target estrogen signaling (Amir et al., 2021) by binding to estrogen receptors or inhibiting the enzyme aromatase, which prevents estradiol synthesis from testosterone. One example of an EDC that effects estrogen signaling is tributyltin (TBT).

Tributyltin is an environmental EDC used as an industrial stabilizer in plastic and PVC pipes, as a fungicide in the timber industry, and as part of antifouling paints used on ship hulls and docks (Hoch, 2001). Extensive use and leaching from antifouling paints resulted in significant levels of TBT in the water column and sediment of aquatic ecosystems. To mitigate these effects, the use of TBT in antifouling paints was internationally banned from use on large boats in 2008 (Gipperth, 2009; IMO, 2019; Showalter and Savarese, 2004 ). Subsequent measurements of environmental TBT levels have decreased (Liang et al., 2017) to the ng/L range (Liu et al., 2020; Zhang, J. et al., 2017) with recovery from imposex observed in New Zealand mollusks (Jones and Ross, 2018). Despite these positive effects, TBT-based paint is still being sold (Uc-Peraza et al., 2022) and increased levels are reported, including ~ 149 ng Sn/g in marinas in coastal areas of Panama (Batista-Andrade et al., 2018), 2 ng/g to more than 500 ng/g in marine protected areas around Latin America (Castro et al., 2021), and ~ 366 ug/kg along the Norway coast (Schoyen et al., 2019). Measurable TBT levels are also found in fish tissue (Qiu et al., 2020; Zhang, J. et al., 2017).

Tributyltin is lipophilic and toxic (Rouleau et al., 2003) with TBT exposure during the early life stages (embryo, larvae, or juveniles) of fish resulting in bioaccumulation (Borges et al., 2014) and a reported high sensitivity of the visual system that includes abnormal and delayed eye development (Hano et al., 2007), thinning of the cornea (Wang and Huang, 1999), and retina-specific defects (de Paulo et al., 2020; Fent and Meier, 1992; Wang and Huang, 1999; Wester et al., 2004). TBT is classically described as an aromatase inhibitor that reduces circulating estradiol levels and increases testosterone levels (Li and Li, 2020), leading to imposex in mollusks (Schoyen et al., 2019) and masculinization in fish (McAllister and Kime, 2003; McGinnis and Crivello, 2011; Santos et al., 2006). These effects occur either by direct inhibition of aromatase or by binding to a PPAR:RXR heterodimer (Ortiz-Villanueva et al., 2018) which then influences aromatase (Lima et al., 2015). In either case, expression of brain aromatase (cyp19a1b) is downregulated following TBT exposure (Lima et al., 2015). Tributyltin also has non-reproductive effects related to oxidative stress (Shi et al., 2021; Zhang, C. et al., 2017), immune responses (Li and Li, 2021; Zhang, C. et al., 2017), lipid accumulation (Barbosa et al., 2019; Ortiz-Villanueva et al., 2018; Zhang, J. et al., 2017), liver function (Zhang et al., 2016a), neurotransmitter synthesis (Liu et al., 2020; Tu et al., 2020; Xiao et al., 2018), and anxiety (Tu et al., 2020; Zhang et al., 2016b).

Zebrafish (Danio rerio), a model animal for toxicological studies, has been used to assess the effects of TBT in both developing and adult animals. Adult exposure increases 11-ketotestosterone and reduces estradiol levels in plasma, consistent with decreased aromatase (cyp19a1b) expression (Lima et al., 2015; Liu et al., 2020). Thyroid hormone levels in zebrafish are also decreased following TBT exposure (Li and Li, 2021). Reproductive changes are noted (Lan et al., 2020; Xiao et al., 2018), as are changes in stress and a variety of behaviors (Liang et al., 2017; Tu et al., 2020). Developmental exposure alters hatching and swimming behaviors (Liang et al., 2017) and reduces growth (Martinez et al., 2019). Most of these studies were performed using chronic (several days or weeks) exposure to TBT. Short term exposure has not been as extensively studied. Given the ability of TBT to be readily taken up and accumulated in fish, we hypothesized that a brief exposure, during development, may have prolonged effects. We specifically focused on the visual system, because of its high sensitivity to TBT (de Paulo et al., 2020; Shi et al., 2021) and the reported TBT-induced neuronal effects (Liu et al., 2020; Tu et al., 2020).

Thus, the purpose of this study was to examine whether transient 24 hour (hr) developmental exposure to TBT has prolonged effects on visual system development and function in zebrafish larvae. We assessed the effects after 1 week of removal from treatment. Zebrafish were exposed at either 24 hpf, 72 hpf, or 7 days (d) pf, ages that correspond to specific events in visual system development. This design allowed us to determine the developmental periods most vulnerable to TBT. Our results show that short-term developmental TBT exposure adversely affected eye development and visually guided optomotor responses and that these effects were persistent, remaining after removal of treatment, and dependent upon age of exposure.

2. Material and methods

2.1. Animal maintenance

Adult zebrafish were maintained at 28 °C on a 14 hr light: 10 hr dark photoperiod in an Aquatic Habitats (Pentair, Apopka, FL) recirculating rack system in the Zebrafish Ecotoxicology, Neuropharmacology, and Vision (ZENV) Laboratory at American University. Larval zebrafish were obtained from in-house spawning of adults throughout September-November 2019. Spawning was accomplished by placing 4–5 fish of each sex into a breeding chamber overnight. The following morning, ~ 1 hr after the beginning of the light photoperiod, fertilized eggs were collected, rinsed, and maintained in petri dishes containing embryo rearing solution (0.006% instant ocean and 0.01% methylene blue) at the same temperature and photoperiod as adult tanks. All protocols were approved by the Institutional Animal Care and Use Committee at American University (protocol #1700 and #20–03).

2.2. Exposures

Exposures occurred when the embryos/larvae were either 24 hpf, 72 hpf, or 7 dpf, as in (Crowley-Perry et al., 2021; Gould et al., 2017). We selected these ages because they correspond to critical periods in the development of the visual system and in the onset of estrogen signaling. Briefly, at 24 hpf the eye has formed, aromatase, and estrogen receptor expression begins (Callard et al., 2001; Mouriec et al., 2009b; Schmitt and Dowling, 1999); at 72 hpf the retina is functional (Schmitt and Dowling, 1999) and larvae have hatched; at 7 dpf visually guided behaviors can be reliably measured and aromatase protein can be detected in retina (Chiang et al., 2001; Le Page et al., 2011; Muto et al., 2005).

At these ages, larvae were transferred to one of four experimental treatments for 24 hr. Tributyltin-chloride (Sigma Chemical Co, St Louis, MO) made as an initial stock solution by dissolving 0.999 μL of 30.72 mM TBT in 499 mL distilled water. Ethanol (EtOH) was the vehicle control. Subsequent experimental solutions were made by diluting the stock solution resulting in experimental concentrations of 0.2 μg /L and 20 μg /L TBT. The vehicle control (0.1% ethanol) used corresponded to the ethanol concentration in the highest TBT exposure group. Thus, each experiment included four treatment groups: vehicle control (EtOH), a system water control (WT; 0.0 μg /L TBT), a low TBT group (0.2 μg/L), and a high TBT group (20 μg/L). These concentrations were chosen because they are within the range of concentrations found in the environment and/or other studies (Amorim et al., 2017; Dong et al., 2006; Lan et al., 2020; Li and Li, 2020; Li and Li, 2021; Liu et al., 2020; Qiu et al., 2020; Showalter and Savarese, 2004; Tu et al., 2020; Zhang, C. et al., 2017). After each exposure period, larvae were transferred to system water and allowed to develop undisturbed for one week. Starting at 3 dpf, fish were fed AP-100 dry food, which was supplemented with Artemia beginning on day 7. Fish were fed daily, prior to water change and debris removal.

To assess survival, zebrafish were counted daily, beginning with the initial exposure and ending at the end of the 7 d post-exposure period when optomotor and startle responses were measured.

2.3. Startle Behavior

A startle test was performed to assess muscular activity/swimming responses. Startle responses were used as a positive control for general swimming ability in the fish, as the visually guided optomotor response requires fish to swim in the direction of the stimulus (see below). Fish were allowed 2 min to acclimate before a 200 g block was dropped from ~57 cm above the counter, about 20 cm from the testing chamber (LeFauve and Connaughton, 2017). Startle responses evoked by dropping the weight were recorded using a Canon FS40 handheld video camera. Video was recorded for 20 sec after dropping the weight to account for any irregularities.

2.4. Optomotor response (OMR)

The OMR stimulus was a rotating, radial, black and white pinwheel (Gould et al., 2017; LeFauve and Connaughton, 2017; LeFauve et al., 2021) generated using PsychoPi software on a Macbook Pro laptop and projected onto a 36 in Dell flat screen computer monitor below the fish. Fish were placed into a 9 cm glass petri dish (6/dish) in ~ 1 cm of water with a 5 mL centrifuge tube (1.9 cm in diameter) placed in the center of the dish to ensure fish were unable to cross the middle of the rotating stimulus. The test chamber was placed directly onto the computer monitor and wrapped in white tape to ensure the only transparent side was the bottom of the tank. The stimulus rotated in one direction (clockwise or counterclockwise) for 30 sec each, with a 30 sec control period of gray light shown in between. The stimulus sequence ran for 4 min and the swimming response of the fish was recorded using a Canon FS40 handheld video camera. A positive response occurred when the fish swam in the direction of the stimulus. Recordings were performed in an isolation chamber to prevent the fish from using other cues.

2.5. Anatomical analysis

At the completion of the behavioral tests, fish were euthanized in a 0.02% tricaine solution and flash frozen. After four days in −20 °C, 5 randomly selected fish from each age and treatment group were thawed and photographed using an Olympus SZX16 stereomicroscope fitted with an Olympus DP72 color camera and CellSens software. Notochord length and eye diameter were measured for each larva using ImageJ. Notochord length was measured from the most anterior part of the head to the posterior portion of the tail. Eye diameter was measured from the most anterior to the most posterior part of the eye. Each measurement was conducted three times and averaged to reduce error.

2.6. Statistical Analysis

The number of larvae surviving each day post-treatment was assessed using either a two-way ANOVA (72 hpf and 7 dpf) or a Friedman’s test (24 hpf). Two-way ANOVAs were also used to assess differences across exposure age (24 hpf, 72 hpf, 7 dpf) and treatment (WT, vehicle, 20 μg/L TBT, 0.2 μg/L TBT) for startle response, OMRs, and anatomical measurements. If significant differences were noted, a Tukey post hoc test was performed. Significance was assessed at an α-level of 0.05. For some data, a one-way ANOVA was also performed at each age. In this case, α-level was adjusted to 0.01 to reduce the chance of Type I error. Statistics were performed using either SPSS (version 26, IBM) or R.

3. Results

Though zebrafish larvae looked qualitatively similar within a given exposure age group (Fig. 1), treatment-dependent differences in overall survival, growth, and optomotor responses were noted.

Figure 1. Representative micrographs of zebrafish larvae from the different exposure groups.

Figure 1.

Open in a new tab

Larvae in the left column (1A-4A) were exposed at 24 hpf (hours postfertilization). Larvae in the middle column (1B-4B) were exposed when they were 72 hpf. Larvae in the right column (1C-4C) were exposed when they were 7 d (days) pf. Exposure lasted for 24 hours, after which time the larvae were transferred to system water for 1 week. Images were taken at the 1 week post exposure time point when behavioral assessments were performed. Treatment groups: WT = system water, EtOH = ethanol vehicle (0.1%), high TBT (20 μg/L), low TBT (0.02 μg/L).

3.1. Survival

An initial decrease in survival was observed in all treatment groups when exposure occurred at 24 hpf, with the largest decrease in the high TBT group (Fig. 2A). At 2 d post-exposure, survival in all treatment groups stabilized for the remainder of the recovery period. Overall, however, the lowest survival (~50%) was observed in the group exposed to 20 μg/L TBT at 24 hpf. The difference in survival across the 24 hpf exposure groups was significant (p<0.001).

Figure 2. Survival.

Figure 2.

Open in a new tab

Percentage of larvae surviving during the 1 week recovery period following acute exposure to a high concentration of TBT (20 μg/L), a low concentration of TBT (0.2 μg/L), water (WT), or vehicle (EtOH, 0.1%). Exposure occurred when zebrafish were aged (A) 24 hpf, (B) 72 hpf, or (C) 7 dpf and the number of fish was determined daily. A significant difference in survival was observed in larvae exposed at 24 hpf and 7 dpf (asterisks). The difference at 24 hpf was due to the reduced survival in the high TBT treatment group. In contrast, the difference in 7 dpf exposed larvae reflected a time-dependent decrease.

Larvae exposed to TBT when they were either 72 hpf (Fig. 2B) or 7 dpf (Fig. 2C) did not show the initial decrease in survival evident in the 24 hpf exposure group. Survival in larvae exposed at 72 hpf was consistently high (>90%) in all treatment groups, and not affected by either treatment (p = 0.366) or recovery time (p = 0.954). In contrast, survival in the 7 dpf exposure groups steadily decreased with time, reaching ~50% on recovery day 6. The significant time-dependent decrease in survival (p = 0.006) was not affected by treatment (p = 0.947), though the lowest overall survival was again in the high TBT treatment group.

3.2. Startle Behavior

Startle responses evoked after the 1 week recovery revealed a significant main effect of treatment (two-way ANOVA, p = 0.008), with responses in the low TBT exposure group significantly reduced compared to both controls. Exposure age also significantly affected these startle responses, with significantly less larvae displaying a startle response when exposure occurred at 24 hpf or 7 dpf, compared to 72 hpf (two-way ANOVA, p = 0.01). This significance is likely driven by the significantly different startle responses measured in the 24 hpf exposure group (one-way ANOVA, p = 0.013; Fig. 3A.). There was no statistically significant interaction between age*treatment (two-way ANOVA, p = 0.49). There were also no differences in startle responses measured in larvae exposed when they were either 72 hpf (p = 0.095) or 7 dpf (p = 0.647).

Figure 3. Startle responses.

Figure 3.

Open in a new tab

Startle responses of larval zebrafish exposed to system water (WT), ethanol (EtOH) vehicle (0.1%), low TBT (0.2 μg/L) or high TBT (20 μg/L) when aged (A) 24 hpf, (B) 72 hpf, and (C) 7 dpf. Data presented are the mean (± SE) percentage of fish responding to a 200 g weight being dropped adjacent to the recording chamber. Startle responses were recorded after 1 week of recovery/removal from treatment. Significant differences are identified with an asterisk, bars with the same letter were not different from each other. The number at the bottom of each bar represents then number of fish measured. P-values are given above each bar. Significance was evaluated at α = 0.05.

3.3. OMR

There was a statistically significant interactive effect between treatment*exposure age on positive OMRs (Fig. 4; two-way ANOVA, p < 0.001) as well as main effects of both exposure age and treatment (p < 0.001). Significant differences were observed between 24 hpf vs. 72 hpf treated larvae and between 24 hpf vs. 7 dpf exposure groups. No difference was noted between larvae exposed 72 hpf and 7 dpf.

Figure 4. Optomotor responses (OMR).

Figure 4.

Open in a new tab

OMRs were assessed after 1 week of recovery/removal from treatment. Exposure ages were (A) 24 hpf, (B) 72 hpf, or (C) 7 dpf. For each graph, the mean (± SE) percentage of fish in each treatment displaying a positive OMR is shown. At all ages, TBT exposure significantly reduced the percentage of larvae displaying a positive OMR (p < 0.001 for all ages). Bars with the same letters are not significantly different from each other. WT = water; EtOH = 0.1% ethanol (vehicle control); high TBT = 20 μg/L; low TBT = 0.2 μg/L. N = 40 for each treatment group. p-values are given above each bar. Significance was evaluated at α = 0.05.

In larvae exposed at 24 hpf, the number of larvae displaying a positive OMR 1 week later differed significantly across all treatment groups (one-way ANOVA; p < 0.001; Fig. 4A). The number of larvae displaying a positive OMR was reduced in both TBT treatment groups, with the greater reduction in the high TBT group. Significant differences due to treatment were also observed in the 72 hpf (Fig. 4B) and 7 dpf (Fig. 4C) exposure groups (p < 0.001 for both). At these older exposure ages, the number of larvae displaying a positive OMR in both control groups were statistically similar and significantly greater than the number of positive responders in either TBT treatment group.

In general, ≤ 50% of TBT-treated larvae displayed positive OMRs, regardless of exposure age. These decreases were observed following 1 week of recovery (i.e., after 1 week removed from treatment) suggesting prolonged effects of exposure.

3.4. Anatomical measurements

There was no significant interaction between the effects of treatment*exposure age on notochord length (Fig. 5; p = 0.233), though significant main effects of age (p < 0.001) and treatment (p < 0.001) were observed. More specifically, age-specific differences were identified in the 24 hpf and 72 dpf exposure groups. When exposure occurred at 24 hpf, larvae in both TBT exposure groups were smaller than the vehicle control (p = 0.001; Fig. 5A). When TBT exposure occurred at 72 hpf (Fig. 5B) larvae in the high TBT group were significantly smaller than the vehicle control (p = 0.002). No significant differences in length were observed when TBT exposure occurred at 7 dpf (p = 0.479; Fig. 5C).

Figure 5. Notochord length.

Figure 5.

Open in a new tab

Mean (± SE) length measurements of larval zebrafish exposed to treatment conditions at either (A) 24 hpf, (B) 72 hpf, or (C) 7 dpf were made after 1 week of recovery. Differences were observed in larvae exposed at 24 hpf and 72 hpf (asterisks) when TBT-treated larvae were smaller than vehicle controls. Bars with the same letters are not significantly different. WT = water; EtOH = 0.1% ethanol (vehicle control); high TBT = 20 μg/L; low TBT = 0.2 μg/L. N = 5 fish were measured for each age and treatment group. p-values are given above each bar. Significance was evaluated at α = 0.05.

Eye diameter measurements displayed significant main effects of both exposure age and treatment (p < 0.001) as well as a significant interactive effect (p < 0.001). The results of the posthoc analysis revealed that overall eye diameter measurements taken from larvae exposed at 24 hpf were significantly different from measurements taken from larvae exposed at both older ages; measurements taken from TBT exposed larvae were significantly different from controls. Looking at each exposure age individually, measurements in larvae exposed to TBT at 24 hpf were significantly smaller compared to the vehicle control (p < 0.001; Fig. 6A). Measurements from larvae in the high TBT group were also significantly different from WT controls. Decreases in eye diameter were observed in larvae treated with high TBT when they were 7 dpf (p = 0.002; Fig. 6C). No significant differences in eye diameter were observed in larvae exposed at 72 hpf (p = 0.86; Fig. 6B).

Figure 6. Eye diameter.

Figure 6.

Open in a new tab

Mean (± SE) eye diameter measurements of larval zebrafish exposed to treatment conditions at either (A) 24 hpf, (B) 72 hpf, or (C) 7 dpf. Significant differences were observed in groups exposed at 24 hpf and 7 dpf (asterisk) when smaller measurements were observed in the high TBT treatment group. WT = water; EtOH = 0.1% ethanol (vehicle control); high TBT = 20 μg/L; low TBT = 0.2 μg/L. N = 5 fish were measured for each age and treatment group. p-values are given above each bar. Significance was evaluated at α = 0.05.

4. Discussion

Our results show that transient developmental exposure to TBT during the first week after fertilization has deleterious effects on zebrafish larvae. These effects are age-dependent, with the most severe effects observed when exposure occurred at 24 hpf (Fig. 7). Our results further suggest sensitivity of the visual system to TBT. Smaller eye diameters were evident when larvae were treated with TBT at either 24 hpf or 7 dpf. Larvae displaying positive optomotor responses were reduced regardless of the age of exposure. These effects were observed 1 week after removal from treatment, suggesting persistent effects. Further, while most effects were noted in the high (20 μg/L) TBT treatment group, they were often also seen in the low (0.2 μg/L) exposure group, indicating TBT-induced effects were not always concentration dependent.

Figure 7. Summary diagram.

Figure 7.

Open in a new tab

Summary diagram comparing the results obtained across exposure ages. When embryos (24 hpf, blue) were exposed to TBT, decreased survival, growth, startle, and optomotor responses were observed. Larvae exposed when they were 72 hpf (red) were smaller with a decreased number displaying a positive OMR. If exposure occurred when larvae were 7 dpf (green), both eye diameter and the number of fish displaying a positive OMR were decreased. These behavioral and anatomical differences were noted after 1 week in recovery conditions, indicating transient exposure to TBT caused persistent effects. The consistent reduction in positive optomotor responses, coupled to decreases in eye size, suggest vision-specific effects.

Endocrine disruptors, such as TBT, are a significant public health concern because exposure is associated with a variety of diseases. Most EDCs target estrogen pathways (Gelinas and Callard, 1993; TEDX, 2011) causing potent effects on developmental and adult reproductive physiology. For example, as an aromatase inhibitor (Matthiessen and Gibbs, 1998; McAllister and Kime, 2003; McGinnis and Crivello, 2011) TBT prevents estrogen synthesis, leading to masculinization in fish (McAllister and Kime, 2003; McGinnis and Crivello, 2011; Santos et al., 2006). TBT also downregulates aromatase B gene expression in brain (Lan et al., 2020; Lima et al., 2015) and alters estrogen receptor expression (Morales et al., 2013). Non-reproductive effects of TBT exposure are varied and include damage to DNA (Morales et al., 2013), altered expression of aryl hydrocarbon receptors (Mortensen and Arukwe, 2007) and reduced thyroid hormone synthesis (Li and Li, 2021; Sharan et al., 2014). Tributyltin exposure also induces oxidative stress (Shi et al., 2021; Zhang, C. et al., 2017), affects neurotransmitter systems (Liu et al., 2020; Tu et al., 2020), and triggers an immune response (Li and Li, 2021; Zhang, C. et al., 2017). These effects are observed following chronic (several weeks of) exposure.

Our experimental paradigm transiently exposed zebrafish aged 24 hpf, 72 hpf, or 7 dpf to TBT for 24 hr. Larvae were then removed from treatment and allowed to develop until they were 9 dpf, 11 dpf, and 15 dpf, respectively. We did not observe any qualitative differences in larvae post-recovery with respect to age or treatment, however, survival differences were noted in the 24 hpf exposure group. Here, the lowest survival was observed in the high TBT treatment group. This finding contrasts previous studies reporting no differences in survival when zebrafish larvae were chronically exposed to TBT from 2 to 5 dpf (Liang et al., 2017; Martinez et al., 2020) or in juvenile medaka exposed for 4 weeks (Shi et al., 2021). However, these studies used a lower concentration of TBT, which likely accounts for survival differences.

We also identified reduced growth in zebrafish exposed to both TBT concentrations at 24 hpf and 72 hpf. However, startle responses, which were used as a control for movement, were only significantly reduced in the 24 hpf exposure group. These results suggest TBT age-dependently decreases growth and swimming behaviors, with the largest effect observed when exposure occurred at 24 hpf. Decreased growth rate and/or swimming behaviors following TBT exposure were previously reported in zebrafish larvae (Liang et al., 2017; Martinez et al., 2019), in 6-day old tropical guppies (de Paulo et al., 2020), and in medaka juveniles (Shi et al., 2021). Adult female zebrafish chronically exposed to TBT since 5 dpf were smaller than exposed males (Lima et al., 2015).

The visual system is particularly sensitive to TBT exposure. Juvenile tiger perch exposed to 10, 50, or 100 μg/L TBT for 7 days displayed thinning of the cornea and a d0se-dependent decrease in RPE retinomotor responses (Wang and Huang, 1999). Similarly, tropical guppies developmentally exposed to ≤ 818 ng/L TBT for 7 days displayed changes to their retinas which included RPE depigmentation, retinomotor movement dysfunction, increased disorganization of the photoreceptor layer, vacuoles in the inner plexiform layer, darker irises, and a thickened cornea (de Paulo et al., 2020). Tributyltin injected into fertilized medaka eggs caused concentration-dependent mortality and morphological defects, including abnormal eye development (Hano et al., 2007). Medaka juveniles orally exposed to TBT for 4 weeks showed increased antioxidant activity in the eye that was correlated with decreased locomotor and social activity (Shi et al., 2021). Zebrafish larvae continually exposed to 20 μg/L TBT until 60–66 hpf displayed a 2–3 fold increase in retinal apoptosis and a 2-fold increase in the number of macrophages in retina; morphological changes were observed at higher TBT levels (Dong et al., 2006). Lower concentrations (0.0003 to 0.3 uM) of TBT (Martinez et al., 2019) decreased eye length in larval zebrafish. Unique to the current study is the finding that a transient 24 hr exposure to TBT was sufficient to reduce eye diameters in treated fish and decrease optomotor motor responses 1 week after the end of the exposure period.

More specifically, we found that TBT exposure at 7 dpf reduced eye diameter and decreased the number of fish displaying a positive optomotor responses. These results are similar to our previous study (Gould et al., 2017) using the clinically-relevant aromatase inhibitor 4-OH-A (formestane). In that study, we employed the same experimental design, i.e., transient developmental exposure at 24 hpf, 72 hpf, and 7 dpf, but we assessed optomotor responses in adult fish after 3–4 months of recovery/removal from treatment. As found here, OMRs were reduced in developmentally-treated adult zebrafish, with exposure age a significant variable (Gould et al., 2017). That study also found reduced eye diameters in adults treated with 4-OH-A when they were 7 dpf. Together with the current work, these studies suggest TBT’s mode of action may be to inhibit aromatase activity, though we can’t discount that TBT may also be triggering a different pathway independent of estrogen synthesis/aromatase activity (Ortiz-Villanueva et al., 2018). For example, TBT exposure could be altering ATPase activity in gills and glucose metabolism in muscle (Li and Li, 2020) or inhibiting genes necessary for the synthesis of specific neurotransmitters (Tu et al., 2020), all of which could affect vision-based behaviors. We are currently performing molecular analyses to identify how TBT is exerting its effects at the cellular level.

Eye morphogenesis in zebrafish begins at 12 hpf with evagination of left and right optic primordia (Schmitt and Dowling, 1994). The primordia undergo structural changes and rotation over the next 12 hr resulting in well-formed optic cups at 24 hpf (Schmitt and Dowling, 1994). At this time, development of the neural retina begins in the inner (vitreal) to outer (scleral) direction. During the next 24 hr (our earliest exposure age), ganglion and amacrine cells form, followed by horizontal and photoreceptor cells; bipolar cells are formed last (Schmitt and Dowling, 1999). After ganglion cell formation, a small number of ganglion cell axons leave the eye and form the optic nerve at ~34–36 hpf (Burrill and Easter, 1995; Schmitt and Dowling, 1994, 1999; Stuermer, 1988). Initial differentiation of photoreceptors occurs in a patch ventral to the optic nerve head at ~2–2.5 dpf (50 hpf) (Kjlavin, 1987; Schmitt and Dowling, 1999). Between 60–70 hpf all neuronal cell types can be identified and synapses become apparent (Schmitt and Dowling, 1999). At hatching (70–74 hpf) – our next treatment age – all cell types are present, the retina appears functional (Schmitt and Dowling, 1999), the optic tectum is innervated (Stuermer, 1988), and visually-guided behaviors occur (Easter and Nicola, 1996, 1997). One week after fertilization (7 dpf – our final exposure age), visually guided OMRs can be recorded (Chiang et al., 2001; Muto et al., 2005).

Estrogen signaling is present during this morphological development. Estrogen receptor expression begins at 24–48 hpf (Bardet et al., 2002; Lassiter et al., 2002; Shi et al., 2013) throughout developing embryos (Tingaud-Sequeira et al., 2004), which corresponds to the onset of aromatase mRNA expression (Mouriec et al., 2009b). Aromatase expression also increases between 24–48 hpf (Kishida and Callard, 2001), slightly after onset of estrogen receptor transcription and at the same time as estrogen responsiveness (Callard et al., 2001 ). Aromatase can be detected in retina using immunocytochemistry at 7 dpf (Le Page et al., 2011). Previous studies have shown that TBT exposure causes abnormal and delayed eye development (Hano et al., 2007), thinning of the cornea (Wang and Huang, 1999), retina-specific defects (Dong et al., 2006; Fent and Meier, 1992; Wang and Huang, 1999; Wester et al., 2004), and long-term deficits in vision-based behaviors (Gould et al., 2017). Our results are consistent with these reports showing sensitivity of the visual system to TBT. Further, our experimental paradigm, which targeted critical periods in the development of both visual and estrogen signaling systems, suggests vision-specific effects of this compound may be due to a combined effect on both systems.

The effects of TBT occur because it is lipophilic (Morabito, 1995), readily bioaccumulates (Hajjah el Hassani et al., 2005; Morabito, 1995; Oliver et al., 2011), and crosses the blood brain barrier (Rouleau et al., 2003). Bioaccumulation has been observed at concentrations comparable to those used here (Borges et al., 2014; Hajjah el Hassani et al., 2005; Oliver et al., 2011), with internal levels increasing with both duration and concentration used (Hajjah el Hassani et al., 2005). Studies with 72 hpf zebrafish identified measurable TBT levels after 45–48 hr of exposure as well as 3 d after removal from treatment (Borges et al., 2014) indicating relatively slow clearance rates. Our results support this conclusion as we observed behavioral and anatomical differences 1 week after removal from treatment across all ages tested.

Developmental aromatase inhibition with 4-OH-A impairs larval swimming behaviors (Hano et al., 2007; Schmidt et al., 2005; Schmidt et al., 2004) due to deficits in spinal motoneuron development (Houser et al., 2011; Nelson et al., 2008). Altered swimming behaviors have been reported in TBT-treated carp (Schmidt et al., 2005; Schmidt et al., 2004) and medaka (Hano et al., 2007). Given that OMRs require the fish to swim in the direction of stimulus movement, it was important to determine if TBT exposure affected locomotion in our study. To determine this, we recorded startle responses in exposed larvae. We found no differences in startle responses in larvae exposed to TBT at either 72 hpf or 7 dpf, though significant and maintained decreases in positive OMR were noted at both ages. Optomotor responses were also significantly reduced after 1 week of recovery in the 24 hpf exposure group. It should be noted, however, that startle responses were also reduced when TBT exposure occurred at 24 hpf, suggesting a general/systemic effect of exposure. In agreement with this, we observed the most consistent effects when exposure occurred at 24 hpf, with all experimental time points altered. In contrast, larvae exposed when they were 72 hpf and 7 dpf displayed survival comparable controls, a significant (50%) reduction in positive OMR, and (for the 7 dpf group) reduced eye diameters. This suggests that TBT caused sublethal effects with differential, age-dependent effects on eye/visual circuitry.

5. Conclusions

Our results identify morphological and behavioral evidence in zebrafish of the persistent effects of transient exposure to tributyltin during development. These findings also hold significant implications in the field of policy and public health. Due to the rising prevalence of environmental contaminants and related morbidities, and the observed spikes in TBT concentration in various coastal locations, more research should be focused on elucidating long-term effects and preventing exposure to these harmful chemicals. Furthermore, aromatase inhibitors, the class of chemicals which includes TBT, are used in the treatment of metastatic, hormone receptor positive breast center in post-menopausal women (Chumsri et al., 2011). This research could inform closer examinations of visual system effects of aromatase inhibitors and potentially alleviating these effects via antagonizing the mechanism in which TBT operates.

HIGHLIGHTS.

  • Zebrafish larvae were transiently exposed to environmental concentrations of TBT.

  • Visually-guided optomotor responses were reduced 1 week postexposure.

  • Eye diameter was reduced in larvae exposed to TBT at 24hpf or 7dpf.

  • Transient developmental TBT exposure has persistent effects on the visual system.

Acknowledgments

The authors would like to thank Mikayla Crowley-Perry for demonstrating experimental methods and creating stock dilutions of chemicals. The authors would also like to thank Angelo Barberio, Erica Winston, and Mikayla Crowley-Perry for assistance with data analysis protocols and animal maintenance.

Declaration of interests

The authors declare the following financial interests/personal relationships which may be considered as potential competing interests:

Victoria P. Connaughton reports financial support was provided by National Institutes of Health.

Funding

This project was funded by NIH Grant R15EY029866–01 (to VPC).

Footnotes

Conflicts of interest:

The authors declare no conflicts of interest.

Publisher's Disclaimer: This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

References

  1. Alonso-Magdalena P, Quesada I, Nadal A, 2011. Endocrine disruptors in the etiology of type 2 diabetes mellitus. Nat Rev Endocrinol 7, 346–353. [DOI] [PubMed] [Google Scholar]
  2. Amir S, Shah S, Mamoulakis C, Docea A, Kalantzi O-I, Zachariou A, Calina D, Carvalho F, Sofikitis N, Makrigiannakis A, Tsatsakis A, 2021. Endocrine disruptors acting on estrogen and androgen pathways cause reproductive disorters through multiple mechanisms: a review. Int J Environ Res Public Health 18, 1464. [DOI] [PMC free article] [PubMed] [Google Scholar]
  3. Amorim J, Fernandes M, Vasconcelos V, Teles L, 2017. Evaluation of the sensitivity spectrum of a video tracking system with zebrafish (Danio rerio) exposed to five different toxicants. Environ Sci Pollut Res 24, 116086–116096. [DOI] [PubMed] [Google Scholar]
  4. Baier H, Klostermann S, Trowe T, Karlstrom R, Nusslein-Volhard C, Bonhoeffer F, 1996. Genetic dissection of the retinotectal projection. Development 123(415–425). [DOI] [PubMed] [Google Scholar]
  5. Barbosa M, Capela R, Rodolfo J, Fonseca E, Montes R, Andre A, Capitao A, Carvalho A, Quintana J, Castro L, Santos M, 2019. Linking chemical exposure to lipid homeostasis: a municipal waste water treatment plant influent is obesogenic for zebrafish larvae. Ecotoxicol Environ Safety 182, 109046. [DOI] [PubMed] [Google Scholar]
  6. Bardet P-L, Horard B, Robinson-Rechavi M, Laudet V, Vanacker J-M, 2002. Characterization of oestrogen receptors in zebrafish (Danio rerio). J Mol Endocrinol 28, 153–163. [DOI] [PubMed] [Google Scholar]
  7. Batista-Andrade J, Caldas S, Batista R, Castro I, Fillmann G, Primel E, 2018. From TBT to booster biocides: levels and impacts of antifouling along coastal areas of Panama. Environ Poll 234, 243–252. [DOI] [PubMed] [Google Scholar]
  8. Borges A, Oliver AL-S, Gallego-Gallegos M, Munoz-Olivas R, Vale M, Camara C, 2014. Transformation of tributyltin in zebrafish eleutheroembryos (Danio rerio). Biol Trace Elem Res. [DOI] [PubMed] [Google Scholar]
  9. Burrill J, Easter S, 1995. The first retinal axons and their microenvironment in zebrafish: cryptic pioneers and the pretract. J Neurosci 15, 2935–2947. [DOI] [PMC free article] [PubMed] [Google Scholar]
  10. Callard G, Tchoudakova A, Kishida M, Wood E, 2001. Differential tissue distribution, developmental programming, estrogen regulation and promoter characteristics of cyp19 genes in teleost fish. J Steriod Biochem Molec Biol 79, 305–314. [DOI] [PubMed] [Google Scholar]
  11. Casals-Casas C, Desvergne D, 2011. Endocrine disruptors: from endocrine to metabolic disruption. Ann Rev Physiol 73, 135–162. [DOI] [PubMed] [Google Scholar]
  12. Castro I, Machado F, de Sousa G, Paz-Villarraga C, Fillmann G, 2021. How protected are marine protected areas: a case study of tributyltin in Latin America. J Environ Manag 278, 111543. [DOI] [PubMed] [Google Scholar]
  13. Chiang E-L, Yan Y-L, Tong S-K, Hsiao P-H, Guigen Y, Postlethwait J, Chung B-C, 2001. Characterization of duplicated zebrafish cyp19 genes. J Exp Zool 290, 709–714. [DOI] [PubMed] [Google Scholar]
  14. Chumsri S, Howes T, Bao T, Sabnis G, Brodie A, 2011. Aromatase, aromatase inhibitors, and breast cancer. J Steriod Biochem Molec Biol 125, 13–22. [DOI] [PMC free article] [PubMed] [Google Scholar]
  15. Crowley-Perry M, Barberio A, Zeino J, Winston E, Connaughton V, 2021. Zebrafish optomotor response and morphology are altered by transient, developmental exposure to bisphenol-A. J Dev Biol 9, 14. [DOI] [PMC free article] [PubMed] [Google Scholar]
  16. de Paulo D, Mariz C Jr, de Melo Alves M, Alves R, Batista R, Fillmann G, Carvalho P, 2020. Histological and behavioral toxicity of tributyltin in the tropical guppy Poecilia vivipara. Environ Toxicol Chem 39, 1953–1963. [DOI] [PubMed] [Google Scholar]
  17. Diamanti-Kandarakis E, Bourguignon J-P, Giudice L, Hauser R, Prins G, Soto A, Zoeller R, Gore A, 2009. Endocrine-disrupting chemicals: an endocrine society scientific statement. Endocr Rev 30, 293–342. [DOI] [PMC free article] [PubMed] [Google Scholar]
  18. Dong W, Muramoto W, Nagai Y, Takehana K, Stegeman J, Teraoka H, Hiraga T, 2006. Retinal neuronal cells is a toxicological target of tributyltin in developing zebrafish. J Vet Med Sci 68, 573–579. [DOI] [PubMed] [Google Scholar]
  19. Easter S, Nicola G, 1996. The development of vision in the zebrafish (Danio rerio). Dev Biol 180, 646–663. [DOI] [PubMed] [Google Scholar]
  20. Easter S, Nicola G, 1997. The development of eye movements in the zebrafish (Danio rerio). Dev Psychobiol 31, 267–276. [DOI] [PubMed] [Google Scholar]
  21. Fent K, Meier W, 1992. Tributyltin-induced effects on early life stages of minnows Phoxinus phoxinus. Arch Environ Contam Toxicol 22, 428–438. [DOI] [PubMed] [Google Scholar]
  22. Gelinas D, Callard G, 1993. Immunocytochemical and biochemical evidence for aromatase in neurons of the retina, optic tectum and retinotectal pathways in goldfish. J Neuroendocrinol 5, 635–641. [DOI] [PubMed] [Google Scholar]
  23. Gipperth L, 2009. The legal design of the international and European Union ban on tributyltin antifouling paing: direct and indirect effects. J Environ Manag 90, S86–S95. [DOI] [PubMed] [Google Scholar]
  24. Gould C, Wiegand J, Connaughton V, 2017. Acute developmetnal exposure to 4-hydroxyandrostenedione has a long term effect on visually-guided behaviors. Neurotoxicol Teratol 64, 45–49. [DOI] [PubMed] [Google Scholar]
  25. Hajjah el Hassani L, Frenich A, Vidal J, Muros M, Benjiba M, 2005. Study of the accumulation of tributyltin and triphenyltin compounds and their main metabolites in the sea bass, Dicentrachus labrax, under laboratory conditions. Sci Total Environ 348, 191–198. [DOI] [PubMed] [Google Scholar]
  26. Hamad A, Kluk M, Fox J, Park M, Turner J, 2007. The effects of aromatase inhibitors and selective estrogen receptor modulators on eye development in the zebrafish (Danio rerio). Curr Eye Res 32, 819–827. [DOI] [PubMed] [Google Scholar]
  27. Hano T, Oshima Y, Kim S, Satone H, Oba Y, Kitano T, Inoue S, Shimasaki Y, Honjo T, 2007. Tributyltin causes abnormal development in embryos of medaka, Oryzias latipes. Chemosphere 69, 927–933. [DOI] [PubMed] [Google Scholar]
  28. Hatch E, Nelson J, Stahlhut R, Webster T, 2010. Association of endocrine disruptors and obesity: perspectives from epidemiologic studies. Int J Androl 33, 324–332. [DOI] [PMC free article] [PubMed] [Google Scholar]
  29. Hoch M, 2001. Organotin compounds in the environment - an overview. Applied Geochem 16, 719–743. [Google Scholar]
  30. Houser A, McNair C, Piccinini R, Luxhoj A, Bell W, Turner J, 2011. Effects of estrogen on the neuromuscular system in the embryonic zebrafish (Danio rerio). Brain Res 1381, 106–116. [DOI] [PubMed] [Google Scholar]
  31. IMO, 2019. International Convention on the Control of Harmful Anti-fouling Systems on Ships. https://www.imo.org. 2021). [DOI] [PubMed]
  32. Jones M, Ross P, 2018. Recovery of the New Zealand muricid dogwhelk Haustrum scobina from TBT-induced imposex. Marine Poll Bull 126, 396–401. [DOI] [PubMed] [Google Scholar]
  33. Kajta M, Wojtowicz A, 2013. Impact of endocrine-disrupting chemicals on neuronal development and the onset of neurological disorders. Pharmacol Reports 65, 1632–1639. [DOI] [PubMed] [Google Scholar]
  34. Karlstrom R, Trowe T, Klostermann S, Baier H, Brand M, Crawford A, Grunewald B, Haffter P, Hoffmann H, Meyer S, Muller B, Richter S, van Eeden F, Nusslein-Volhard C, Bonhoeffer F, 1996. Zebrafish mutations affecting retinotectal axon pathfinding. Development 123, 427–438. [DOI] [PubMed] [Google Scholar]
  35. Kishida M, Callard G, 2001. Distinct cytochrome P450 aromatase isoforms in zebrafish (Danio rerio) brain and ovary are differentially programmed and estrogen regulated during early development. Endocrinology 142, 740–750. [DOI] [PubMed] [Google Scholar]
  36. Kjlavin I, 1987. Early development of photoreceptors in the ventral retina of the zebrafish embryo. J Comp Neurol 260, 461–471. [DOI] [PubMed] [Google Scholar]
  37. Lan X-R, Li Y-W, Chen Q-L, Shen Y-J, Liu Z-H, 2020. Tributyltin impaired spermatogenesis and reproductive behavior in male zebrafish. Aquat Toxicol 224, 105503. [DOI] [PubMed] [Google Scholar]
  38. Lassiter C, Kelley B, Linney E, 2002. Genomic structure and embryonic expression of estrogen receptor beta a (ERBa) in zebrafish (Danio rerio). Gene 299, 141–151. [DOI] [PubMed] [Google Scholar]
  39. Le Page Y, Vosges M, Servili A, Brion F, Kah O, 2011. Neuroendocrine effects of endocrine disruptors in teleost fish. J Toxicol Env Health B 14, 370–386. [DOI] [PubMed] [Google Scholar]
  40. LeFauve M, Connaughton V, 2017. Impacts of heavy metal exposure alters visually-guided behaviors in zebrafish. Curr Zool 63, 221–227. [DOI] [PMC free article] [PubMed] [Google Scholar]
  41. LeFauve M, Gould C, Crowley-Perry M, Wiegand J, Shapiro A, Connaughton V, 2021. Using a variant of the optomotor response as a visual defect detection assay in zebrafish. J Biol Meth 8, e144. [DOI] [PMC free article] [PubMed] [Google Scholar]
  42. Li P, Li Z-H, 2020. Tributyltin induces the tissue-specific stresses in zebrafish, a study in various tissues of muscle, gill and intestine. Bull Environ Contamin Toxicol 105, 847–852. [DOI] [PubMed] [Google Scholar]
  43. Li Z-H, Li P, 2021. Effects of the trybutyltin on the blood parameters, immune responses and thyroid hormone system in zebrafish. Environ Pollution 268, 115707. [DOI] [PubMed] [Google Scholar]
  44. Liang X, Souders C III, Zhang J, Martyniuk C, 2017. Tributyltin induces premature hatching and reduces locomotor activity in zebrafish (Danio rerio) embryos/larvae at environmentally relevant levels. Chemosphere 189, 498–506. [DOI] [PubMed] [Google Scholar]
  45. Lima D, Castro L, Coelho I, Lacerda R, Gesto M, Soares J, Andre A, Capela R, Torres T, Carvalho A, Santos M, 2015. Effects of tributyltin and other retinoid receptor agonists in reproductive-related endpoints in the zebrafish (Danio rerio). J Toxicol Env Health A 78, 747–760. [DOI] [PubMed] [Google Scholar]
  46. Liu Z-H, Li Y-W, Hu W, Chen Q-L, Shen Y-J, 2020. Mechanisms involved in tributyltin-enhanced aggressive behaviors and fear responses in male zebrafish. Aquat Toxicol 220, 105408. [DOI] [PubMed] [Google Scholar]
  47. Macdonald R, Wilson S, 1996. Pax proteins and eye development. Curr Opin Neurobiol 6, 49–56. [DOI] [PubMed] [Google Scholar]
  48. Martinez R, Codina A, Barata C, Tauler R, Pina B, Navarro-Martin L, 2020. Transcriptomic effects of tributyltin (TBT) in zebrafish eleutheroembryos. A functional benchmark dose analysis. J Hazard Mat 398, 122881. [DOI] [PubMed] [Google Scholar]
  49. Martinez R, Herrero-Nogareda L, Van Antro M, Campos M, Casado M, Barata C, Pina B, Navarro-Martin L, 2019. Morphometric signatures of exposure to endocrine disrupting chemicals in zebrafish eleutheroembryos. Aquat Toxicol 214, 105232. [DOI] [PubMed] [Google Scholar]
  50. Matthiessen P, Gibbs P, 1998. Critical appraisal of the evidence for tributyltin-mediated endocrine disruption in mollusks. Environ Toxicol Chem 17, 37–43. [Google Scholar]
  51. McAllister B, Kime D, 2003. Early life exposure to environmental levels of the aromatase inhibitor tributyltin causes masculinisation and irreversible sperm damage in zebrafish (Danio rerio). Aquat Toxicol 65, 309–316. [DOI] [PubMed] [Google Scholar]
  52. McGinnis C, Crivello J, 2011. Elucidating the mechanisms of action of tributyltin (TBT) in zebrafish. Aquat Toxicol 103, 25–31. [DOI] [PubMed] [Google Scholar]
  53. Morabito R, 1995. Speciation of organotin compounds in environmental matrices. Microchem J 51, 198–206. [Google Scholar]
  54. Morales M, Martinez-Paz P, Ozaez I, Martinex-Guitarte J, Morcillo G, 2013. DNA damage and transcriptional changes induced by tributyltin (TBT) after short in vivo exposures of Chironomus riparius (Diptera) larvae. Comp Biochem Physiol C 158, 57–63. [DOI] [PubMed] [Google Scholar]
  55. Mortensen A, Arukwe A, 2007. Modulation of xenobiotic biotransformation system and hormonal responses in Atlantic salmon (Salmo salar) after exposure to tributyltin (TBT). Comp Biochem Physiol C 145, 431–441. [DOI] [PubMed] [Google Scholar]
  56. Mouriec K, Lareyre J, Tong S-K, Le Page Y, Vaillant C, Pelligrini E, Pakdel F, Chung B-C, Kah O, Anglade I, 2009b. Early regulation of brain aromatase (cyp19a1b) by estrogen receptors during zebrafish development. Dev Dyn 238, 2641–2651. [DOI] [PubMed] [Google Scholar]
  57. Muto A, Orger M, Wehman A, Smear M, Kay J, Page-McCaw P, Gahtan E, Xiao T, Nevin L, Gosse N, Staub W, Finger-Baier K, Baier H, 2005. Forward genetic analysis of visual behavior in zebrafish. PLoS Genet 1, e66. [DOI] [PMC free article] [PubMed] [Google Scholar]
  58. Nelson B, Henriet R, Holt A, Bopp K, Houser A, Allgood O Jr, Turner J, 2008. The role of estrogen in developmental appearance of sensory-motor behaviors in the zebrafish (Danio rerio): the characterization of the “listless” model. Brain Res 1222, 118–128. [DOI] [PubMed] [Google Scholar]
  59. Oliver AL-S, Sanz-Landaluze J, Munoz-Olivas R, Guinea J, Camara C, 2011. Zebrafish larvae as a model for the evaluation of inorganic arsenic and tributyltin bioconcentration. Water Res 45, 6515–6524. [DOI] [PubMed] [Google Scholar]
  60. Ortiz-Villanueva E, Jaumot J, Martinez R, Navarro-Martin L, Pina B, Tauler R, 2018. Assessment of endocrine disruptors effects on zebrafish (Danio rerio) embryos by untargeted LC-HRMS metabolomic analysis. Sci Total Environ 635, 156–166. [DOI] [PubMed] [Google Scholar]
  61. Qiu X, Takamura T, Enoki S, Kato-Unoki Y, Takai Y, Nagano Y, Kinoshita M, Kitano T, Shimasaki Y, Oshima Y, 2020. Detoxification roles of tributyltin-binding protein type 2 in Japanese medaka (Oryzias latipes) exposed to tributyltin. Mar Poll Bull 159, 111445. [DOI] [PubMed] [Google Scholar]
  62. Rouleau C, Xiong Z-H, Pacepavicius G, Huang G-L, 2003. Uptake of waterborne tributyltin in the brain of fish: axonal transport as the proposed mechanism. Environ Sci Technol 37, 3298–3302. [DOI] [PubMed] [Google Scholar]
  63. Santos M, Micael J, Carvalho A, Morabito R, Booy P, Massanisso P, Lamoree M, Reis-Henriques M, 2006. Estrogens counteract the masculinizing effect of tributyltin in zebrafish. Comp Biochem Physiol C 142, 151–155. [DOI] [PubMed] [Google Scholar]
  64. Schmidt K, Staaks G, Pflugmacher S, Steinberg C, 2005. Impact of PCB mixture (Aroclor 1254) and TBT and a mixture of both on swimming behavior and body growth ane enzymatic biotransformation activities (GST) of young carp (Cyprinis carpio). Aquat Toxicol 71, 49–59. [DOI] [PubMed] [Google Scholar]
  65. Schmidt K, Steinberg C, Pflugmacher S, Staaks G, 2004. Xenobiotic substances such as PCB mixtures (Aroclor 1254) and TBT can influence swimming behavior and biotransformation activity (GST) of carp (Cyprinis carpio). Environ Toxicol 19, 460–470. [DOI] [PubMed] [Google Scholar]
  66. Schmitt E, Dowling J, 1994. Early eye morphogenesis in the zebrafish, Brachydanio rerio. J Comp Neurol 344, 532–542. [DOI] [PubMed] [Google Scholar]
  67. Schmitt E, Dowling J, 1999. Early retinal development in the zebrafish, Danio rerio: ligth and electron microscopic analysis. J Comp Neurol 404, 515–536. [PubMed] [Google Scholar]
  68. Schoyen M, Green N, Hjermann D, Tveiten L, Beylich B, Oxnevad S, Beyer J, 2019. Levels and trends of tributyltin (TBT) and imposex in dogwhelk (Nucella lapillus) along the Norwegian coastline from 1991 to 2017. Mar Environ Res 144, 1–8. [DOI] [PubMed] [Google Scholar]
  69. Schug T, Blawas A, Gray K, Heindel J, Lawler C, 2015. Eludicating the links between endocrine disruptors and neurodevelopment. Endocrinology 156, 1941–1951. [DOI] [PMC free article] [PubMed] [Google Scholar]
  70. Sharan S, Nikhil K, Roy P, 2014. Disruption of thyroid hormone functions by low dose exposure of tributyltin: an in vitro and in vivo approach. Gen Comp Endocrinol 206(155–165). [DOI] [PubMed] [Google Scholar]
  71. Shi Y, Chen C, Li M, Liu L, Dong K, Chen K, Qiu X, 2021. Oral exposure to tributyltin induced behavioral abnormalities and oxidative stress in the eyes and brains of juvenile Japanese medaka (Oryzias latipes). Antioxidants 10, 1647. [DOI] [PMC free article] [PubMed] [Google Scholar]
  72. Shi Y, Liu X, Zhu P, Li J, Sham K, Cheng S, Li S, Zhang Y, Cheng C, Lin H, 2013. G-protein-coupled estrogen receptor1 is involved in brain development during zebrafish (Danio rerio) embryogenesis. Biochem Biophys Res Comm 435, 21–27. [DOI] [PubMed] [Google Scholar]
  73. Showalter S, Savarese J, 2004. Restrictions on the use of marine antifouling paints containing tributyltin and copper. p. 11.
  74. Stuermer C, 1988. Retinotopic organizaiton of the developing retinotectal projection in the zebrafish embryo. J Neurosci 8, 4513–4530. [DOI] [PMC free article] [PubMed] [Google Scholar]
  75. TEDX, 2011. The TEDX list of potential endocrine disruptors. https://endocrinedisruption.org/interactive-tools/tedx-list-of-potential-endocrine-disruptors/search-the-tedx-list. (Accessed 6/4/2013.
  76. Tingaud-Sequeira A, Andre M, Forgue J, Barthe C, Babin P, 2004. Expression patterns of three estrogen receptor genes during zebrafish (Danio rerio) development: evidence for high expression in neuromasts. Gene Expression Patterns 4, 561–568. [DOI] [PubMed] [Google Scholar]
  77. Trowe T, Klostermann S, Baier H, Granato M, Crawford A, Grunewald B, Hoffmann H, Karlstrom R, Meyer S, Muller B, Richter S, Nusslein-Volhard C, Bonhoeffer F, 1996. Mutations disrupting the ordering and topographic mapping of axons in the retinotectal projection of the zebrafish, Danio rerio. Development 123, 439–450. [DOI] [PubMed] [Google Scholar]
  78. Tu X, Li Y-W, Chen Q-L, Shen Y-J, Liu Z-H, 2020. Tributyltin enhanced anxiety of adult male zebrafish through elevating cortisol level and disruption of serotonin, dopamine and gamma-aminobutyric acid neurotransmitter pathways. Ecotoxicol Environ Safety 203, 111014. [DOI] [PubMed] [Google Scholar]
  79. Uc-Peraza R, Castro I, Fillmann G, 2022. An absurd scenario in 2021: banned TBT-based antifouling products still available on the market. Sci Total Environ 805, 150377. [DOI] [PubMed] [Google Scholar]
  80. Wang W-Y, Huang B-Q, 1999. TBT (tributyltin) toxicity to the visual and olfactory functions of tigerperch (Terapon jarbua Forsskal). Zool Studies 38, 189–195. [Google Scholar]
  81. Wester P, van der Ven L, Vos J, 2004. Comparative toxicological pathology in mammals and fish: some examples with endocrine disruptors. Toxicol 205, 27–32. [DOI] [PubMed] [Google Scholar]
  82. Xiao W-Y, Li Y-W, Chen Q-L, Liu Z-H, 2018. Tributyltin impaired reproductive success in female zebrafish through disrupting oogenesis, reproductive behaviors and serotonin synthesis. Aquat Toxicol 200, 206–216. [DOI] [PubMed] [Google Scholar]
  83. Zhang C, Zhang J, Ren H, Zhou B, Wu Q, Sun P, 2017. Effect of tributyltin on antioxidant ability and immune responses of zebrafish (Danio rerio). Ecotoxicol Environ Safety 138, 1–8. [DOI] [PubMed] [Google Scholar]
  84. Zhang J, Sun P, Kong T, Yang F, Guan W, 2016a. Tributyltin promoted hepatic steatosis in zebrafish (Danio rerio) and the molecular pathogenesis involved. Aquat Toxicol 170, 208–215. [DOI] [PubMed] [Google Scholar]
  85. Zhang J, Zhang C, Sun P, Huang M, Fan M, Liu M, 2017. RNA-sequencing and pathway analysis reveal alteration of hepatic steroid biosynthesis and retinol metabolism by tributyltin exposure in male rare minnow (Gobiocypris rarus). Aquat Toxicol 188, 109–118. [DOI] [PubMed] [Google Scholar]
  86. Zhang J, Zhang C, Sun P, Shao X, 2016b. Tributyltin affects shoaling and anxiety behavior in female rare minnow (Gobiocypris rarus). Aquat Toxicol 178, 80–87. [DOI] [PubMed] [Google Scholar]

RESOURCES