Diquat based herbicide impair the development and behavior of zebrafish embryos and larvae

Abstract

The growing pressure on food production has led to the widespread use of pesticides to ensure high crop yields. Herbicides are extensively applied in Brazil, with diquat-based herbicides (DBH) being one of the most commercialized. DBH has gained prominence as a substitute for paraquat and as a solution for glyphosate-resistant weeds. However, strict regulations govern the permissible levels of these substances in drinking water. In this context, we evaluated environmentally relevant concentrations of DBH in zebrafish (Danio rerio), a model organism in ecotoxicological studies. To evaluate the impact on behavior and early development, embryos were exposed to DBH at concentrations of 20, 50, or 250 µg L⁻¹ during organogenesis (3–120 hpf). Developmental impacts were assessed through survival rates, spontaneous movement, and heart rate, while behavioral effects were analyzed using open-field and aversive stimulus tests. DBH exposure reduced survival, increased spontaneous movements and heart rate, and impaired larval behavior. Notably, all tested concentrations induced adverse effects in at least one endpoint, clearly demonstrating the detrimental impact of DBH on zebrafish embryos and larvae at environmentally relevant concentrations.

Supplementary Information

The online version contains supplementary material available at 10.1038/s41598-025-26040-x.

Introduction

Aquatic fauna faces a constant and complex reality, as water resources are the recipients of many environmental contaminants¹. In addition, there is intense pressure for food production driven by food insecurity. To achieve high productivity, the use of agricultural inputs from different classes and toxicities is necessary.

In recent years, Brazil has stood out globally for its large-scale commercialization of pesticides², coupled with a significant increase in the registration of new substances. Therefore, the presence of pesticides in the country’s water resources is inevitable³. However, this reality is not exclusive to Brazil; worldwide, the presence of pesticides is well-documented⁴, leading regulatory agencies to adopt measures to curb their presence in aquatic environments.

Since 2009, the Brazilian Institute of the Environment and Renewable Natural Resources (IBAMA) has issued reports on pesticide sales in the country, allowing us to understand the agribusiness profile and the most widely used substances with the highest potential to reach the environment. In recent years, one compound has drawn attention: diquat dibromide, which ranked 10th in the latest available report (2023)⁵.

Diquat-based herbicide (DBH) are gaining traction in the Brazilian market, as they are used on pre-harvest desiccation⁶, weed control^7,8, and aquatic macrophytes⁹. Moreover, the increased use of DBH in Brazil is linked to the 2020 ban on paraquat^10,11. Both compounds share a similar mechanism of action (they inhibit Photosystem I and belong to the bipyridyl chemical group), which act as false electron acceptors in Photosystem I, leading to ATP synthesis dropdown and to increased production of reactive oxygen species, ultimately causing plant death through the oxidative degradation of fatty acids in thylakoid and other cellular membranes via free radical formation¹². Under Brazilian conditions, the half-life of diquat was 29 days in the water–plant–soil system, 23 days in the water–soil system, and 18 days in water alone¹³.

Given their multiple uses and the known toxicity of pesticides in general, specific regulations exist that establish permissible concentrations for human consumption and drinking water. In Brazil, Ordinance GM/MS No. 888 (2021) regulates water quality control and monitoring for human consumption, setting permissible limits such as 13 µg L⁻¹ for paraquat, 30 µg L⁻¹ for 2,4-D, 180 µg L⁻¹ for tebuconazole, and 500 µg L⁻¹ for glyphosate¹⁴. However, no reference value exists for DBH. Other countries, such as the U.S. and Canada, have established regulatory limits of 20 µg L⁻¹¹⁵ and 50 µg L⁻¹¹⁶ for DBH respectively. Nonetheless, although these are considered ‘safe’ values, it is crucial to investigate potential effects on non-target organisms.

DBH is a highly interesting compound for ecotoxicological studies because, as mentioned above, it is also applied directly to water to control invasive plants, inadvertently exposing non-target species such as fish to contamination. In this context, the early life stages of fish represent a critical window for understanding the effects of chronic pollutant exposure, particularly since hatching and developing in contaminated environments poses major survival challenges for these organisms¹⁷. Along these lines, some species have numerous advantages, such as the zebrafish (Danio rerio), whose embryos are transparent, exhibit rapid development, high prolificity¹⁸, and heightened sensitivity to environmental contaminants. Furthermore, disturbances during the embryonic stage can directly impact animal life in the long term^19–21, making this exposure window highly relevant for ecotoxicological studies. Indeed, zebrafish have been widely used in scientific research, particularly in pesticide studies (e.g.^22,23). Furthermore, the results obtained with zebrafish can be extrapolated to other species due to their broad genetic homology (including > 70% similarity with humans²⁴), highlighting their flexibility as an experimental model for investigations across different exposure scenarios—particularly for toxicity mechanisms conserved among species, such as those examined here (survival, physiology, and behavior).

Few studies have investigated the toxicity of DBH in zebrafish. In adults, DBH caused nuclear abnormality, disturbance of antioxidant system and dysfunction in the liver²², impaired growth and caused hepatotoxicity²³. The effects of this herbicide were also evaluated during this critical developmental window of embryogenesis. In this phase, active ingredient (a.i.) Diquat dibromide exposure caused a specific effect on embryonic movement²⁵, pericardial edema, elevated heart rate and hypochromia²⁶, in addition to disrupting mitochondrial bioenergetics and behavior²⁷, and induced oxidative stress²⁸. Because DBH induces oxidative stress and affects mitochondrial function, endpoints such as heart rate and exploratory behavior serve as integrative readouts of sublethal neurotoxic and physiological impairment. Although some evidence suggests the toxic potential of diquat, existing studies have limitations regarding the risks of the commercial formulation in aquatic ecosystems. This highlights the need to further advance our understanding of this pesticide’s toxicity, especially considering its direct application in aquatic environments for control aquatic weed (e.g.^29,30) and the lack of behavioral outcomes and assessments.

This context highlights a critical research gap: a comprehensive evaluation of a commercial diquat-based herbicide (DBH) throughout the early developmental window of zebrafish, encompassing both physiological and behavioral outcomes. Therefore, our study was designed to directly address these shortcomings by investigating the effects of environmentally relevant concentrations of DBH, based on international drinking water standards, throughout organogenesis (3–120 hpf). We achieved this by integrating developmental measures (survival, spontaneous movement, and heart rate) with behavioral analyses (exploratory and antipredator responses) to provide a more complete understanding of its ecotoxicological impact.

Material and methods

Study strategy

To evaluate the effects of DBH on zebrafish early development, embryos were exposed to environmentally relevant concentrations (20, 50, or 250 µg L⁻¹) or to a negative control during the organogenesis phase from 3 to 120 h post-fertilization (hpf), followed by the evaluation of initial developmental parameters, exploratory behavior, and antipredatory responses.

Reproduction and maintenance of embryos

For reproduction, healthy wild-type (5D strain) zebrafish (aged 6–18 months) were used, commercially acquired and maintained at the Laboratory of Ecology and Conservation of the Federal University of Fronteira Sul. Breeders were fed twice daily with commercial flaked food (SUPERVIT®; Tropical, Chorzow, Poland) raw protein 48%, fat extract, and crude oils 8%), switched off under artificial photoperiod (12 h light: 12 h dark), and with water quality settings according to specifications for the species, which include pH 7.0 ± 0.2; dissolved oxygen at 6.2 ± 0.4 mg L⁻¹ and total ammonia < 0.01 mg L⁻¹ and temperature at 28 ± 2 °C. In the afternoon, breeding pairs (2:1 female-to-male ratio) were separated and placed in dedicated spawning tanks with mesh bottoms to prevent egg predation. The following morning, 1 h after lights have turned on, embryos were collected by siphoning and rinsed with E3 medium (reverse osmosis water + 60 mg L⁻¹ Ocean Tech Bio Active®, Hong Kong, China³¹) to remove debris and feces. Embryos were then washed, sorted, and maintained in 24-well cell culture plates (3 mL well⁻¹), with 10 embryos per well, and incubated in a water bath at 28 °C until 7 days post-fertilization (dpf). Embryos were kept in E3 medium³² under an artificial photoperiod (12 h light: 12 h dark).

Pesticide and concentration tested

The first and second DBH concentrations selected for this study were permitted for human consumption in the USA (20 µg L⁻¹¹⁵) and Canada (50 µg L⁻¹¹⁶). The third concentration of 250 µg L⁻¹ was based on the label rate and was similar to the approximate a.i. concentration in a treatment area³³.

For exposures, we used the commercial product which contains 200 g L⁻¹ diquat and 970 g L⁻¹ other ingredients in its formulation, here named as DBH. Serial solutions were prepared based on 200 g L⁻¹ of active ingredient. First, a 1 g L⁻¹ solution was prepared. Next, a stock solution with a concentration of 0.01 g L⁻¹ was made. From this stock solution, the remaining solutions for the different experimental groups were derived. The prepared herbicide solutions were stored in sealed containers at 4 °C in the refrigerator and protected from light. Solutions were used within the 8 days of preparation, during which their stability and biological activity were maintained³⁴. All solutions were prepared in E3 medium and served as the negative control treatment and all solutions tested had neutral pH (6.9–7.1). We did not perform analytical determination of concentrations in this study, as HPLC analyses confirmed the nominal concentrations—e.g., 1.13 mg L⁻¹ of diquat a.i. from a nominal 1 mg L⁻¹ solution³⁵. The lack of analytical confirmation of the actual concentrations represents a limitation of the present study. However, the combination of product specifications, regulatory assurances, and methodological consistency with previous studies provides sufficient confidence that the toxicological effects reported here are environmentally valid and relevant.

Embryos were exposed during the organogenesis period (3–120 hpf³⁶). Embryos in the control group were maintained in E3 medium alone, whereas treatment groups were exposed to DBH solutions prepared using E3 medium as the base. On the 3rd exposure day (72 hpf), after all animals had hatched, the chorions were removed and the solution was completely renewed. Additionally, dead animals were removed daily and well volumes were replenished when necessary. On 5 dpf (120 hpf), larvae were removed from exposure, rinsed, and transferred to new plates containing only E3 medium for continuation of the experimental protocol. Daily renewal of the concentration was not necessary, as the half-life of diquat is greater than 5 days¹³. For all analyses, animals were randomly selected without reuse across evaluations.

Developmental and survival parameters

Survival

To assess larval survival, fish were examined every morning through 7 dpf. Embryos lacking transparency, exhibiting coagulation, absent cellular formation, or lacking heartbeat and blood circulation were considered dead and removed. For this analysis, 180 embryos per group were monitored (n = 180).

Spontaneous movement and heart rate

At 24 hpf, embryos exhibit spontaneous movements (SM) while still within the chorion, which are induced by motor neuron development without any central nervous system control^18,37,38. For movement counting, the animals were not removed from the exposure plates. The plate was positioned directly under the stereomicroscope (Olympus SZ51) SM were quantified for 60 s³⁷ per embryo (n = 20).

At 72 hpf, we counted the heart rate (HR) for 60 s³⁹ in 20 larvae per group (n = 20) using a binocular biological microscope (Olympus CX21FS1).

Behavioral analysis

Open field test (OFT)

To determine whether DBH altered larval exploratory behavior, at 144 hpf larvae were individually placed in 6-well cell culture plates (total capacity: 15 mL) filled with 10 mL of E3 medium and recorded for 6 min. Thirty larvae per group were evaluated from 6 replicates (n = 30). We used the camera function of an iPhone 11 (iOS 18), and videos were analyzed using the EthoVision XT 17.5 software. The testing arena was virtually divided into a central zone and a peripheral/wall zone. The parameters used to assess larval behavior included total distance (mm), total time moving (s), absolute turn angle (°), latency to first time in the center (s), distance moved in the center (mm), total time in the center (s), distance moved in the periphery (mm), and total time in the periphery (s). This test was based on³¹.

Aversive stimulus test (AST)

The larvae’s antipredatory response was assessed using the aversive stimulus test (AST). For this purpose, 168 hpf larvae were placed in 6-well cell culture plates (total capacity: 15 mL) filled with 10 mL of E3 medium, at a density of 5 larvae per well with 15 replicates (75 larvae per group). The plates were then positioned on an LCD monitor, and after a 2-min acclimation period, we initiated exposure to the visual stimulus—a 1.35 cm red sphere with a trajectory covering only half of the well. The sphere’s movement was generated using PowerPoint software (Microsoft Office Professional Plus 2016), with larvae exposed to this aversive stimulus for 5 min. The test outcome measured the percentage of larvae remaining in the stimulus area. All recordings were made using the camera function of an iPhone 11 (iOS 18). This test was based on^31,40.

Statistics

Sample sizes were determined a priori using a power analysis conducted in G*Power software (α = 0.05, power = 0.8, effect size ≥ 0.5). Survival analysis was performed using the Kaplan–Meier method, and differences among groups were evaluated using the log-rank (Mantel–Cox) test. For HR, SM, OFT, and AST data, we first assessed homoscedasticity and normality through Kolmogorov–Smirnov and Brown-Forsythe tests. Data were then evaluated by one-way ANOVA followed by Dunnett’s post-hoc test (parametric data) or Kruskal–Wallis test followed by Dunn’s post-hoc test (non-parametric data). All groups were compared to the negative control group. The alpha level was set at 0.05. All statistical analyses were performed using GraphPad Prism® version 8.01 (GraphPad Software, San Diego, USA). For more details on the analyses, see the supplementary material (SM).

Results

Survival

Survival analysis using the Kaplan–Meier method revealed significant differences in survival among treatments (χ² = 44.87, df = 3, p < 0.0001; log-rank test).

Pairwise log-rank (Mantel–Cox) comparisons showed that all exposure groups had significantly lower survival compared with the control. Specifically, survival in the 20 µg L⁻¹ group was reduced relative to the control (χ² = 8.87, df = 1, p = 0.0029), with a hazard ratio (HR) of 0.37 [95% CI 0.20–0.68], indicating a 63% lower risk of survival compared with control. The 50 µg L⁻¹ group exhibited a more pronounced reduction (χ² = 35.27, df = 1, p < 0.0001) with an HR of 0.18 [95% CI 0.11–0.29], corresponding to a 5.5-fold higher mortality risk relative to control. Similarly, the 250 µg L⁻¹ group showed a marked decrease in survival (χ² = 34.48, df = 1, p < 0.0001) with an HR of 0.18 [95% CI 0.11–0.29].

These results demonstrate a clear concentration-dependent decrease in survival probability across treatments. Kaplan–Meier survival curves for all groups are presented in Fig. 1, Tables S1, and S2.

Fig. 1.

Kaplan–Meier survival curves of zebrafish larvae exposed to different concentrations of a diquat-based herbicide (DBH). Data represent the percentage of surviving individuals over time. Survival distributions were compared using the log-rank (Mantel–Cox) test following Kaplan–Meier estimation. Asterisks denote significant differences in survival relative to the control group (**p < 0.001; ****p < 0.0001; N = 180).

Spontaneous movement (SM) and heart rate (HR)

Exposure to DBH increased SM at concentrations of 20 µg L⁻¹ and 50 µg L⁻¹ (H = 8.620; p = 0.0348; Fig. 2A, Table S3) and increased HR at 50 µg L⁻¹ and 250 µg L⁻¹ (F_{(3, 76)} = 8.592; p < 0.0001; Fig. 2B, Table S3).

Fig. 2.

Spontaneous movement (SM) (A) and heart rate (HR) (B) of zebrafish embryos exposed to different concentrations of diquat-based herbicide (DBH). (A) SM was analyzed by Kruskal–Wallis test followed by Dunn’s test and data are expressed as the mean ± SEM. HR was analyzed by one-way ANOVA followed by Dunnet’s multiple comparison test and data are expressed as mean ± SEM. Asterisks indicate statistical differences about the control group (*p < 0.05; ***p < 0.001; ****p < 0.0001; N = 20).

Open field test (OFT)

Exposure to DBH decreased the total distance (p = 0.01; Fig. 3A), total time moving (p < 0.0001; Fig. 3B), distance moved in the periphery (p = 0.01; Fig. 3G), but did not alter absolute turn angle (p > 0.05; Fig. 3C), latency to first time in the center (p > 0.05; Fig. 3D), distance moved in the center (p > 0.05; Fig. 3E), total time in the center (p > 0.05; Fig. 3F), and total time in the periphery (p > 0.05; Fig. 3H) at a concentration of 20 µg L⁻¹. The other concentrations evaluated did not change the behavior of the embryos in comparison to control. More details are available in Table S3. As no trend test was performed to statistically evaluate the dose–response pattern, the relationship between concentration and effect could not be definitively characterized.

Fig. 3.

Open field test results for fish exposed to different concentrations of diquat-based herbicide (DBH). (A) Total distance (mm), (B) total time moving (s), (C) absolute turn angle, (D) latency to first time in the center (s), (E) distance moved in the center (mm), (F) total time in the center (s), (G) distance moved in the periphery (mm), (H) total time in the periphery (s). The data was compared using a One-way ANOVA followed by Tukey test blocked by replicate and are expressed as mean ± S.E.M. (*p < 0.05, ****p < 0.0001).

Aversive stimulus test (AST)

Exposure to DBH did not alter the anti-predatory response at any of the concentrations tested. The data are available in Table S3.

Discussion

Here we demonstrate that DBH exposure affected survival, increased heart rate and spontaneous movements, and impaired zebrafish larvae behavior. Indeed, all tested concentrations elicited effects in at least one evaluated endpoint, clearly showing the impact of DBH exposure on zebrafish embryos and larvae.

Exposure to all tested DBH concentrations (20, 50, or 250 µg L⁻¹) compromised animal survival, consistent with effects observed in another fish species (Oncorhynchus mykiss) embryos and larvae also exposed to DBH⁴¹. In other studies, no significant mortality was observed for zebrafish larvae exposed to diquat, however, only the active ingredients (a.i.) was used^27,28. Several studies have reported that the product formulations exhibit higher toxicity compared to their a.i. alone⁴². Commercial formulations contain co-formulants, which are considered trade secrets and therefore are not disclosed in the product label⁴³; it remains unclear whether they have their own toxic effects or if they enhance the toxicity of the a.i.⁴⁴. The mortality and other observed effects in this study may be associated with the presence of co-formulants in the DBH. However, testing the commercial formulation is more relevant to our study, as it is the actual product applied in the environment, where it comes into contact with water and may impact aquatic organisms. Nevertheless, despite this context, we recognize that evaluating individual components would be valuable for determining their specific contributions to the observed toxicity. This approach could include testing both the purified active ingredient and isolated excipients. Future studies in this direction could help clarify the underlying mechanisms of the effects we observed, particularly regarding potential synergistic interactions among formulation components. Furthermore, we highlight that DBH is sold in Argentina, Mexico, Peru, Uruguay, Chile, Bolivia, Colombia, Costa Rica, Cuba, Canada, United States, El Salvador, Guatemala, Honduras, Haiti, Nicaragua, Panama, Paraguay, Guyana, China, among other countries⁴⁵. This widespread use underscores the urgency of understanding its ecological impact, particularly given the uncertainties surrounding co-formulant toxicity.

We observed lethal effects in zebrafish embryos and larvae even at these concentrations considered safe’ for human consumption (20 and 50 µg L⁻¹ of the diquat) and concentration in a treatment area (250 µg L⁻¹; approximate value, calculated based on the product’s application rate³³). In other words, two of these concentrations are legally permitted in some countries, and the highest one could be found in an area following herbicide application. These mortality increases (> 5.5-fold) observed here indicate potential population-level consequences under repeated or widespread exposures; however, extrapolation to field populations requires data on environmental concentrations and exposure frequency. Similar effects are well-documented for other pesticides including glyphosate⁴⁶, diflubenzuron, pyriproxyfen and their mixtures⁴⁷, chlorpyrifos and p,p′-DDE⁴⁸ atrazine and diuron⁴⁹, and atrazine⁵⁰ underscoring the importance of assessing pesticide contamination impacts during developmentally sensitive stages.

Another significant finding was the increased heart rate observed at 50 and 250 µg L⁻¹ DBH exposure, indicating potential cardiotoxicity. As one of the first organs to be formed in fish, the heart is fundamentally linked to early development⁵¹. These cardiac alterations, combined with the changes in SM, suggest DBH induces early developmental disruptions. Cardiac acceleration in zebrafish has also been documented following exposure to pyriproxyfen⁴⁷, carbaryl and fenitrothion⁵², and fenpropathrin⁵³. Multiple mechanisms may explain this dysfunction: (1) Oxidative stress²⁸ and mitochondrial impairment²⁷ may trigger compensatory heart rate adjustments to maintain cardiac output; (2) Direct stress responses to pesticide exposure, as heart rate serves as a sensitive, immediate stress indicator in fish⁵⁴.

Embryos exposed to concentrations of 20 and 50 µg L⁻¹ showed increased SM. This analysis was performed at 24 hpf, when animals remain within the chorion, demonstrating that DBH penetrated the semipermeable protective barrier of the chorion and interfered with involuntary movements. SM, which emerge during early development, are mediated by intrinsic neural activity⁵⁵, these results indicate neurotoxic interference occurred in DBH-exposed animals. The ability of compounds to cross the chorion and disrupt neuromotor processes during early developmental stages reinforces the disruptive potential of xenobiotics, as shown here and also in⁵⁶. Furthermore, SM alterations are recognized as sensitive toxicity markers in embryo assays⁵⁷, potentially predicting morphological or behavioral disorders in later developmental stages. No gross malformations were observed in any treatment group, which supports that the observed effects on SM were not confounded by severe developmental abnormalities.

Continuing this behavioral analysis, at 144 hpf, DBH-exposed animals showed reduced exploratory activity, with decreased total distance moved, reduced time moving, and less periphery exploration at 20 µg L⁻¹. Interestingly, this contrasts with reported DBH-induced hyperactivity in zebrafish larvae exposed for 7 days to 10–100 μM, but using a.i. diquat²⁷. Independent of the product type and variations in responses, diquat induced behavioral changes. These findings reinforce that intact behavioral patterns are crucial for organismal fitness⁵⁸. The observed alterations may significantly impact survival, reproduction, and ecological balance^59,60, potentially increasing predator vulnerability while impairing foraging and reproductive success. Although our study did not detect significant alterations in the antipredator response assessed by the AST, we cannot rule out the possibility that such effects may occur under different experimental conditions (e.g., other concentrations, formulations, or exposure durations). This interpretation is supported by studies demonstrating impairment of this response in fish exposed to both pesticides⁶¹ and pharmaceuticals³¹. Nevertheless, it is also worth mentioning that the standardized group-based nature of this test may mask more subtle individual-level impairments, including those related to social interactions between larvae.

In our study, exploratory behavior was significantly altered only at the lowest concentration tested (20 µg L⁻¹). We note that no formal statistical trend test was performed; thus, the following interpretation remains a plausible hypothesis. The absence of effects at higher concentrations, coupled with significant effects at the lowest tested dose, suggests a potential non-monotonic dose–response relationship, a phenomenon often observed in behavioral toxicology. Such a pattern could be explained by hormetic mechanisms or receptor-mediated thresholds, where low-dose exposures trigger stimulatory or disruptive effects that are not observed at higher doses⁶². Alternatively, it is possible that we observed behavioral overstimulation at the low dose versus a generalized sedative effect or systemic toxicity at higher concentrations, which may have masked specific exploratory changes by suppressing overall activity. Although less likely given the strength of the effects observed, the possibility of a chance finding due to multiple comparisons must also be considered. While further mechanistic studies are needed to confirm these possibilities, our findings highlight the importance of considering non-linear responses when evaluating behavioral outcomes of contaminant exposure.

Over the long term, such effects—survival, cardiotoxicity and changes in SM and behavior—may severely compromise species perpetuation by reducing population sizes and directly impact population dynamics. This finding raises important concerns since DBH is used for aquatic weed control⁹, meaning aquatic organisms face continuous exposure to varying concentrations of this compound, consequences that remain poorly investigated, particularly regarding survival impacts. Most alarmingly, the two lowest concentrations tested here (20 and 50 µg L⁻¹) fall within drinking water limits permitted in several countries^15,16.

We did not evaluate the possible mechanisms involved in the effects observed here. Additionally, the reported DBH concentrations are nominal, since no analytical verification was performed on the exposure media. However, this does not diminish the merit of our findings, given the scarcity of studies assessing the effects of DBH on fish—particularly during the embryonic stage and regarding behavioral aspects. Nevertheless, reports in the literature suggest that the redox cycle and consequent ROS generation are cytotoxic mechanisms induced by bipyridyl herbicides⁶³, in addition to oxidative stress^22,27,28, bioenergetic disruption, and mitochondrial dysfunction²⁷.

A plausible mechanism is oxidative stress via superoxide generation; this hypothesis is supported by prior studies showing redox cycling and ROS formation⁶³, mitochondrial dysfunction²⁷, and biomarkers of oxidative stress after diquat exposure^27,64,65. Finally, these processes contribute to oxidative stress, inflammation, and the activation of cell death mechanisms such as apoptosis or necrosis, depending on the severity and location of the damage. Therefore, we hypothesize that these cytotoxic effects are closely associated with the observed behavioral impairments, particularly reduced spontaneous movement and altered habituation rate, which serve as sensitive indicators of neurotoxicity.

An intriguing observation, however, is that alterations in exploratory behavior were only evident at the lowest diquat concentration. While increased oxidative stress typically induces upregulation of antioxidant enzymes, recent findings in other species indicate that although SOD transcript levels were elevated in 22 µg L⁻¹ diquat-exposed Lymnaea stagnalis, this did not consistently translate into higher enzymatic activity⁶⁵. This suggests that diquat may impair processes downstream of transcription, potentially affecting translation, protein folding, or enzyme stability—a mechanism that, while not investigated in the present study, could help contextualize our behavioral results. Consequently, even in the presence of transcriptional upregulation, insufficient SOD activity would allow superoxide to persist and exert toxic effects. At lower concentrations, where antioxidant defenses may not yet be fully activated or are functionally impaired despite transcriptional cues, subtle yet significant behavioral disruptions can manifest. In contrast, higher concentrations may elicit a stronger compensatory response or induce more severe toxicity that masks specific behavioral changes. These findings emphasize the complexity of oxidative stress responses and highlight how discrepancies between gene expression and functional enzyme activity can influence toxicological outcomes, particularly in neurobehavioral endpoints.

Furthermore, we cannot rule out the potential persistence of the observed effects, especially since similar exposure windows (0–120 h) in other contaminant classes have shown that animals carried embryonic-stage contamination effects into adulthood and even transmitted them to subsequent generations^19,21,66. Additionally, the same concentrations tested in this study were previously reported as toxic to the amphibian species Physalaemus cuvieri, causing reduced survival, increased body mass index (BMI) and scaled mass index (SMI), malformations, and alterations in swimming behavior⁶⁷. These findings raised significant concerns about the compound’s potential adverse effects and motivated the present study to investigate whether such effects would recur in another aquatic organism—a hypothesis confirmed by our results. Zebrafish exposed to DBH similarly exhibited reduced survival and behavioral alterations, demonstrating DBH’s potential to induce significant changes in non-target aquatic species. This comparison becomes interesting, since the amphibians were in their aquatic larval stage, a phase ecologically comparable to fish, as both require water for survival and are directly exposed to contaminants.

Our findings are environmentally relevant and reinforce the importance of the United Nations Sustainable Development Goals (SDGs), which aim to ensure a balanced environment for future generations. This research aligns with SDG 6 (Clean Water and Sanitation) and SDG 14 (Life Below Water), demonstrating through significant effects on a non-target species that more initiatives must be implemented to achieve these goals while genuinely preserving the environment.

Conclusion

DBH exposure affected survival, increased spontaneous movements and heart rate, and impaired zebrafish larvae behavior. Our findings demonstrate that even concentrations permitted by law for drinking water (20 and 50 µg L⁻¹) cause effects on aquatic organisms, highlighting the need for further investigations to support revisions of norms and legislation regulating water approved for human consumption, aiming at One Health protection.

Furthermore, we emphasize the need for additional studies investigating the effects of these legally permitted concentrations on potential mechanisms of action (including oxidative stress biomarkers), different exposure windows (chronic and life-cycle scenarios), and possible persistent and transgenerational effects, particularly given the possible occurrence of these concentrations in natural aquatic environments where these organisms live and reproduce. These comprehensive approaches are essential to verify whether currently established ‘safe’ levels truly pose no risk to aquatic biota, and our findings combined with future evidence can support more precise regulatory decisions to enhance environmental protection.

Supplementary Information

Below is the link to the electronic supplementary material.

Author contributions

A.P. Conceptualization, methodology, writing—original draft, visualization. F.B.C: Methodology. W.A.T, J.L.F.: Writing—review and editing. P.A.H, M.H.: Conceptualization, methodology, writing—review and editing, supervision.

Data availability

The datasets used and/or analysed during the current study available from the corresponding author on reasonable request.

Declarations

Competing interests

The authors declare no competing interests.

Ethical approval

All applicable international, national, and/or institutional guidelines for the care and use of animals were followed. This study complies with the guidelines of the National Council for Animal Experimentation Control (CONCEA) and was approved by the Ethics Commission for Animal Use Committee (CEUA) of the Federal University of Fronteira Sul, Erechim, RS, Brazil (Protocol 7408041024). This article does not contain studies with human participants performed by any the authors. The study design, conduct, and reporting adhered to the ARRIVE guidelines to ensure rigorous and reproducible research.