Abstract
Chlorantraniliprole (CHL) is the most widely used diamide worldwide, with South America being its primary market. Despite its growing application, the environmental effects of CHL on nontarget organisms, mainly native fish species, remain understudied. In this study, the sublethal effects of CHL were assessed in Cnesterodon decemmaculatus by acute exposure (96 hr) to 1/10 (1.5 mg/L) and 1/100 (0.15 mg/L) of the median lethal concentration, using a multi-biomarker approach across different levels of biological organization. Locomotor activity (distance traveled, time immobile, average, and maximum speeds), somatic index, enzymatic activities of acetyl-cholinesterase (AChE) in muscle and brain, catalase (CAT) in muscle, brain, gills, and liver, glutathione-S-transferase (GST) in gills and liver, aspartate amino-transferase (AST), alanine amino-transferase (ALT), AST to ALT ratio, and alkaline phosphatase in the liver were measured. The primary effect of exposure was the reduction in locomotor activity, which appears to be more closely related to CHL’s mode of action than cholinergic effects. The muscles and brain were the organs most affected by oxidative stress, and adaptive responses involving AChE, CAT, and GST were observed, highlighting the organism’s ability to manage oxidative stress. The Integrated Biomarker Response (IBR) index indicates a dose-dependent relationship, with individuals exposed to T2 exhibiting more than twice the IBR value of those exposed to T1 and nearly four times that of the control group. Our results indicate that insect-specific compounds like diamides can severely affect nontarget species, potentially affecting survival and growth rates in aquatic species, even at sublethal concentrations. For muscle-targeted insecticides, locomotor activity is one of the most effective biomarkers for assessing the impact of exposure. This study represents the first report on the toxicity of a diamide in a native South American model fish, a key bioindicator in assessing ecological health.
• Effects of acute sublethal exposition chlorantraniliprole (CHL) were evaluated in Cnesterodon decemmaculatus
• Behavioral, morphological, and biochemical biomarkers were applied.
• Locomotor impairment was the most significant toxic effect.
• The Integrated Biomarker Response index indicates disturbance levels increase with concentration.
• CHL may have significant consequences at individual and ecological levels.
Introduction
Insecticides are extensively used in agriculture and public health to control pests that threaten crops, livestock, or human health (Araújo et al., 2023; Rezende-Teixeira et al., 2022). Commonly used insecticides, including carbamates, pyrethroids, neonicotinoids, and organophosphates, are neurotoxic substances that interfere with the normal functioning of the nervous system in target and nontarget organisms (Ansari et al., 2014; Araújo et al., 2023; Arya et al., 2022; Pérez-Parada et al., 2018; Rezende-Teixeira et al., 2022).
Recently, diamides have emerged as promising alternatives to older classes of insecticides. Their action involves selective binding to the ryanodine receptor (RyR), a highly specific biochemical pathway in insects. The resulting uncontrolled release of calcium leads to paralysis and death of the target insects (Du & Fu, 2023; Li et al., 2023). Furthermore, their broad-spectrum efficacy against diverse insect pests—including species from the orders Lepidoptera, Coleoptera, Diptera, and Hemiptera—combined with a relatively low rate of resistance development, has established them as a valuable tool in pest management strategies across various crop systems (Du & Fu, 2023; Sgarbi et al., 2023).
Chlorantraniliprole (CHL) is currently the most extensively utilized diamide insecticide worldwide (Li et al., 2023). Since 2020, South America has led worldwide pesticide application (Food and Agriculture Organization of the United Nations, 2024) and has become the largest market for CHL, accounting for approximately 35% of the global demand (Global Overview: Insecticides Market, 2024). Their extensive use has led to the increased occurrence of this pesticide in freshwater bodies across several South American countries (Alana Dos Santos Campos et al., 2024; Francelino et al., 2023; Navarro et al., 2024; Peresin et al., 2023; Rodríguez-Bolaña et al., 2023).
Introduced as an alternative for organophosphates and pyrethroids in various agricultural applications (Rezende-Teixeira et al., 2022), CHL exhibits low water solubility (Sw = 0.88 mg/L at 20 °C) and a moderate octanol-water partition coefficient (KOW = 2.86). Despite its low water solubility, it remains fat-soluble, highly persistent in soil (half-life [DT50] = 597 days), and exhibits slow degradation in water (DT50 = 170 days; Pesticide Properties DataBase [PPDB], 2024).
Due to its toxicological characteristics, CHL has been included in the Highly Hazardous Pesticides list compiled by the International Pesticide Action Network (PAN, 2021). Recent studies reported that CHL is highly toxic to soil bacterial communities (Wu et al., 2021) and beneficial to terrestrial insects (Li et al., 2024; Sgarbi et al., 2023; Tuelher et al., 2017). Additionally, a recent study that evaluates the Ecological Risk Assessment using both deterministic and probabilistic approaches found a moderate to high ecological risk for CHL to aquatic biota (Rodríguez-Bolaña et al., 2024). This result is concordant with others that show effects in aquatic ecosystems; it induces oxidative stress in invertebrates (Sgarbi et al., 2023; Wang et al., 2022) and fish (AlMisherfi et al., 2023; Meng et al., 2022; Mohamed et al., 2022; Stinson et al., 2022). Despite these findings, there is a lack of toxicological evaluations of diamides in native aquatic organisms in South America.
Aquatic ecosystems can be contaminated by legacy and current-use insecticides through runoff, leaching, and atmospheric deposition (Latif et al., 2023). Their toxicity can vary depending on factors such as dosage, exposure duration, and specific insecticide formulation. However, even at low doses, they can accumulate in aquatic organisms, affecting them at different biological levels of organization, including the molecular, cellular, tissue, organ, population, community, and ecosystem levels (Bertrand & Iturburu, 2023; Bonifacio et al., 2016; Gonçalves et al., 2021; Sharma et al., 2019; Van der Oost et al., 2003).
Fish are among the most suitable organisms for toxicological evaluation, with several advantages, including shorter lifespans, faster reproductive cycles, and easy maintenance under laboratory conditions (Bertrand & Iturburu, 2023; Pérez-Parada et al., 2018; Slaninova et al., 2009). Pesticides can induce oxidative stress in organisms, resulting from an imbalance between the production of reactive oxygen species (ROS) and the antioxidant defense mechanisms of fish, potentially leading to cellular or organismal damage or death (Gonçalves et al., 2021; Slaninova et al., 2009). The initial responses occur at the molecular level. The most studied biomarkers at this level are related to enzymes involved in biotransformation and antioxidant defenses, such as catalase (CAT), superoxide dismutase, glutathione-S-transferases (GSTs), and glutathione peroxidases (Gonçalves et al., 2021; Van der Oost et al., 2003). Other relevant enzymes for assessing pesticide contamination effects include cholinesterase, commonly used to evaluate neurotoxic effects (Bernal-Rey et al., 2020; Fulton & Key, 2001; Gonçalves et al., 2021), and transaminases, which serve as indicators of hepatic damage in fish (Bonifacio et al., 2016, 2020; De la Torre et al., 1999).
Integrating a multi-biomarker approach across different organizational levels provides a comprehensive framework for assessing the effects of pesticides on aquatic organisms (Mittelbach et al., 2014; Sharma et al., 2019). This approach is essential for understanding stress responses and physiological damage, linking individual-level effects to potential ecological consequences at the population scale (Gonçalves et al., 2021; Van der Oost et al., 2003). In this context, behavioral parameters—including swimming velocity, distance traveled, and immobility—serve as effective biomarkers for evaluating the impact of environmental pollutants on aquatic ecosystems (Bonifacio et al., 2020; Porras-Rivera et al., 2024; Scott & Sloman, 2004; Sharma et al., 2019). Given that diamides target RyR in muscle cells, assessing these parameters is particularly relevant in the case of CHL, as it may significantly affect fish locomotor activity. Detecting significant changes in these biomarkers can provide critical insights into physiological stress and potential damage at individual and population levels (Porras-Rivera et al., 2024). To investigate the potential impact of CHL on aquatic ecosystems, we evaluated the effects of sublethal CHL concentrations on Cnesterodon decemmaculatus by analyzing multiple biomarkers, ranging from molecular to behavioral levels.
Materials and methods
Adult C. decemmaculatus females were selected for the experiments because of the characteristics that make them suitable for toxicity assessments. These include small size, wide distribution, high abundance in South American Pampean basins, and ease of collection and maintenance under laboratory conditions (Pautasso et al., 2023). They have been employed in several studies to evaluate the toxicity of several pesticides (Bertrand & Iturburu, 2023; Bonifacio et al., 2017, 2020; Pautasso et al., 2023; Ruiz de Arcaute et al., 2023) being the predominant species used for bioassays in the region (Bertrand & Iturburu, 2023) and field monitoring (Vidal et al., 2018).
Individuals were collected from the Yuspe River (648320 W; 318170S; Córdoba, Argentina) using a 1 mm mesh dip net. This river serves as a reference site for environmental studies because it is an unpolluted site (Rautenberg et al., 2022; Zambrano et al., 2023). Fish were transported to the laboratory and acclimated for 15 days in 120 L tanks in a temperature-controlled room at 21 ± 1 °C with a 12:12-hr light:dark cycle. The individuals were fed twice daily with commercial fish food throughout this period. All procedures comply with the Guide for Care and Use of Laboratory Animals (National Institutes of Health, 2011).
Acute toxicity test
Short-term (96-hr) static toxicity tests were performed to evaluate the toxicity of CHL (>99%) purchased from HPC Standards GmbH (Cunnersdorf, Germany). Stock solutions were prepared in methanol (MeOH) supplied by J.T. Baker (Darmstadt, Germany) and diluted with dechlorinated tap water to reach the work concentration. Previous studies estimated that the 96 hr median lethal concentration (LC50) of CHL for Channa punctatus and Cirrhinus mrigala were 14.424 mg/L and 16.465 mg/L, respectively (Bantu & Vakita, 2013; Rathnamma & Nagaraju, 2014). Based on these values, fish of C. decemmaculatus were exposed to sub-lethal doses at the following concentrations for 96 hr: 0 mg/L (Control), 0.15 mg/L (T1 = 1/100 of the 96 hr LC50), and 1.5 mg/L (T2 = 1/10 of the 96 hr LC50). Ten individuals per treatment were randomly assigned to 2 L aerated glass aquaria. All bioassays were performed in duplicate. Individuals were not fed during the experiment, and the water in each aquarium was renewed every 24 hr. Fish without respiratory movements and no response to tactile stimuli were considered dead and removed immediately.
The stability of CHL in aqueous environments is well-documented, with studies under controlled laboratory conditions reporting a DT50 ranging from 10 to 150 days (Nogueira et al., 2024; PPDB, 2024; Redman et al., 2020). To assess stability and confirm accurate dosing, the concentrations in aquarium water were measured at 0 and 24 hr using the liquid–liquid extraction method described by Rodríguez-Bolaña et al. (2023).
Behavioral parameters
After the exposure period, each fish was individually transferred to a recording aquarium (25 cm width × 9 cm depth × 25 cm height) containing 2 L of dechlorinated tap water. Once the fish settled at the bottom, their activity was recorded for 10 min. To assess whether CHL affects locomotor activity, we determined the mean speed of mobile episodes (m/s), maximum speed (m/s), duration of immobility (s), and distance traveled (m) for each individual. All parameters were obtained at the end of each trial from video recordings using Animal Tracker, an ImageJ-based tracking application programming interface (Gulyás et al., 2016)
Somatic indexes
Fish were euthanized by severing the spinal cord behind the operculum and subsequently dissected. Standard length (mm) and body weight (g) were measured. The Fulton condition factor (K) was calculated for each fish using the following equation (Froese, 2006):
where W is the body weight (g), and L is the standard length (cm).
To measure enzyme activity, gills, muscles, liver, and brain were removed from individuals in each treatment and control group.
Each liver was weighed to calculate the hepatosomatic index (HSI) using the following equation (Chellappa et al., 1995):
where Wl is the weight of the liver (g), and W is the body weight.
Enzyme activities
Enzyme extracts were prepared from individual fish following Bonifacio et al., (2017). For acetylcholinesterase (AChE), CAT, and GST activities, organs were homogenized in 0.1 M potassium phosphate buffer (pH 6.5) containing 20% glycerol, 1 mM ethylenediaminetetraacetic acid, and 1.4 mM dithioerythritol using a glass homogenizer (Potter-Elvehjem). Samples were centrifuged at 6,900 g and 4 °C for 10 min to separate cell debris from the supernatant.
Acetylcholinesterase activity was measured in brain and muscle homogenates following the method of Ellman et al. (1961). Briefly, this technique is based on the degradation of acetylcholine by AChE into acetate and thiocholine. The latter reacts with dithiobisnitrobenzoic acid, producing a yellow compound with a peak absorbance at 412 nm (ε = 1.36 × 104 M−1 cm−1). Absorbance was recorded every 15 s for 3 min.
Catalase activity was determined according to Beutler (1982) by measuring the decrease in absorbance at 240 nm due to substrate (H2O2) consumption. Following Tagliaferro et al. (2018), activity was measured over 2 min.
Glutathione-S-transferase activity was determined in the liver and gills using 1-chloro-2,4-dinitrobenzene (CDNB) as a substrate at 340 nm, following Habig et al. (1974). Briefly, CDNB conjugates with glutathione, forming a thioester that exhibits absorbance at 340 nm. Absorbance was measured for 3 min at 25 °C (Tagliaferro et al., 2018).
For transaminase and alkaline phosphatase (ALP) activities, livers from individual fish were homogenized in phosphate buffer (pH 7.4) and centrifuged at 15,000 g at 4 °C for 10 min to separate cell debris from the supernatant. Aspartate aminotransferase (AST; l-aspartate-2-oxaloglutarate aminotransferase) and alanine aminotransferase (ALT; l-alanine-2-oxaloglutarate aminotransferase) activities were determined following Reitman and Frankel (1957). The reaction mixture consisted of 2 mmol/L α-ketoglutarate and specific substrates for AST (100 mmol/L aspartate) and ALT (200 mmol/L alanine) in 100 mM phosphate buffer (pH 7.4). The reaction was initiated by adding aliquots of the liver supernatant. After 30 min of incubation, 2,4-dinitrophenylhydrazine reagent was added, and the resulting-colored product was quantified by spectrophotometry at 505 nm. The AST to ALT activity ratio was calculated for each treatment.
Alkaline phosphatase activity (orthophosphoric monoester phosphohydrolase) was determined colorimetrically using a commercial kit (Wiener Lab 1361003; Bonifacio et al., 2017). Enzymatic activity was normalized to the protein content of the sample (Bradford, 1976) and expressed as nkat mg−1 protein.
Integrated Biomarker Response index
Results of AChE, CAT, GST, transaminases, and behavioral parameters were analyzed using the Integrated Biomarker Response (IBR) index, following the method described by Beliaeff and Burgeot (2002) and modified by Devin et al. (2014) to integrate all these biomarkers for assessing the health status of individuals exposed to CHL. The IBR index was calculated using the CALculate IBR Interface (Calibri, https://shiny.otelo.univ-lorraine.fr/calibri/R/) developed by the Laboratory for Continental Environments at Lorraine University (France).
Statistical analysis
Data distributions were analyzed using the Shapiro–Wilk test (Sokal & Rohlf, 1999). Analysis of variance (ANOVA) was performed to compare biological parameters among treatments, followed by Tukey’s post hoc test. The Kruskal–Wallis test was used for nonparametric data, followed by the Dunn test (Sokal & Rohlf, 1999). The experimental unit was an individual fish, with three groups (control, treatment 1, and treatment 2), each containing 10 individuals. The degrees of freedom were 2 (between groups) and 27 (within groups). Differences were considered significant at p < 0.05. Statistical analyses were performed using Paleontological Statistics (PAST) software Ver. 4.13 (Hammer et al., 2001).
Results
Analytic concentrations of CHL
The CHL concentration in water from the control groups was below the detection limits at 0 and 24 hr. In contrast, the lowest and highest concentrations in the treatment groups decreased by 5.31% and 2.01%, respectively, after 24 hr (Table 1).
Concentration of chlorantraniliprole in exposure media at 0 and 24 hr (n = 3).
| Chlorantraniliprole (mg/L) | ||
|---|---|---|
| T0 | T24 | |
| Control | <DL | <DL |
| 0.15 mg/L | 0.158 ± 0.008 | 0.1496 ± 0.004 |
| 1.5 mg/L | 1.842 ± 0.003 | 1.805 ± 0.021 |
| Chlorantraniliprole (mg/L) | ||
|---|---|---|
| T0 | T24 | |
| Control | <DL | <DL |
| 0.15 mg/L | 0.158 ± 0.008 | 0.1496 ± 0.004 |
| 1.5 mg/L | 1.842 ± 0.003 | 1.805 ± 0.021 |
Note. DL = detection limits.
Concentration of chlorantraniliprole in exposure media at 0 and 24 hr (n = 3).
| Chlorantraniliprole (mg/L) | ||
|---|---|---|
| T0 | T24 | |
| Control | <DL | <DL |
| 0.15 mg/L | 0.158 ± 0.008 | 0.1496 ± 0.004 |
| 1.5 mg/L | 1.842 ± 0.003 | 1.805 ± 0.021 |
| Chlorantraniliprole (mg/L) | ||
|---|---|---|
| T0 | T24 | |
| Control | <DL | <DL |
| 0.15 mg/L | 0.158 ± 0.008 | 0.1496 ± 0.004 |
| 1.5 mg/L | 1.842 ± 0.003 | 1.805 ± 0.021 |
Note. DL = detection limits.
These low reductions align with CHLs expected slow degradation rate and demonstrate that, with daily renewals, the concentrations in the treatment groups remained relatively stable throughout the experiment.
Behavioral parameters
The locomotor activity of individuals in the control and treatment groups is shown in Figure 1. The results indicate that exposure to CHL significantly reduced the distance traveled, average speed, and maximum speed while significantly increasing the time immobile of the individuals (ANOVA, p < 0.05). However, the responses varied depending on the parameter analyzed and the exposure concentration. For instance, significant differences in mean distance traveled were observed among all treatments, with the lowest activity recorded at the highest concentration (Tukey post hoc test, C vs. T1, p < 0.01; C vs. T2, p < 0.001; T1 vs. T2, p < 0.05). A dose-dependent response was detected regarding maximum speed, although it was only significant in individuals exposed to the highest CHL concentration (Tukey post hoc test, C vs. T1, p = 0.11; C vs. T2, p < 0.05; T1 vs. T2, p = 0.09).
Behavioral parameters of locomotor activity recorded in Cnesterodon decemmaculatus exposed to chlorantraniliprole. C = control group, T1 = 0.15 mg/L, T2 = 1.5 mg/L. Different letters indicate significant differences among treatments and between treatments and the control group (p < 0.05).
In contrast, for average speed (Tukey post hoc test, C vs. T1, p < 0.05; C vs. T2, p < 0.05; T1 vs. T2, p = 0.07) and time immobile (Tukey post hoc test, C vs. T1, p < 0.05; C vs. T2, p < 0.05; T1 vs. T2, p = 0.31), no dose-dependent response was observed, as significant differences were only found between the control group and the treatment groups.
Somatic indexes
The death of one individual was recorded in the T2 group. No mortality was recorded in either the control group or the T1 group.
After exposure to CHL, the Fulton Condition Factor (K) of individuals and the HSI showed no significant differences among treatments and the control group (F = 0.16, p = 0.89, and F = 0.28, p = 0.56, respectively; Table 2).
Fulton condition factor (K) and hepatic somatic index (HSI) of Cnesterodon decemmaculatus exposed to chlorantraniliprole (n = 29).
| Treatment | K | HSI |
|---|---|---|
| Control | 1.77 ± 0.21 | 1.76 ± 0.25 |
| T1 | 2.41 ± 0.28 | 0.94 ± 0.25 |
| T2 | 1.82 ± 0.11 | 1.39 ± 0.39 |
| Treatment | K | HSI |
|---|---|---|
| Control | 1.77 ± 0.21 | 1.76 ± 0.25 |
| T1 | 2.41 ± 0.28 | 0.94 ± 0.25 |
| T2 | 1.82 ± 0.11 | 1.39 ± 0.39 |
Note. The values are expressed as means ± standard error. T1 = 0.15 mg/L; T2 = 1.5 mg/L.
Fulton condition factor (K) and hepatic somatic index (HSI) of Cnesterodon decemmaculatus exposed to chlorantraniliprole (n = 29).
| Treatment | K | HSI |
|---|---|---|
| Control | 1.77 ± 0.21 | 1.76 ± 0.25 |
| T1 | 2.41 ± 0.28 | 0.94 ± 0.25 |
| T2 | 1.82 ± 0.11 | 1.39 ± 0.39 |
| Treatment | K | HSI |
|---|---|---|
| Control | 1.77 ± 0.21 | 1.76 ± 0.25 |
| T1 | 2.41 ± 0.28 | 0.94 ± 0.25 |
| T2 | 1.82 ± 0.11 | 1.39 ± 0.39 |
Note. The values are expressed as means ± standard error. T1 = 0.15 mg/L; T2 = 1.5 mg/L.
Enzyme activities
Enzyme activity responses to CHL exposure are summarized in Table 3. The results reveal a complex relationship between concentration and biological responses, characterized by a combination of dose-dependent effects, inhibitory/stimulatory responses at low concentrations, and tissue-specific variability.
Enzyme activities (nkat/mg prot) in different tissues of Cnesterodon decemmaculatus exposed to chlorantraniliprole.
| Treatment | Control | T1 | T2 | |
|---|---|---|---|---|
| Biomarker | Organ | |||
| CAT | Gills | 5.94 ± 0.91(ab) | 2.56 ± 0.51(a) | 8.32 ± 1.92(b) |
| Liver | 74.74 ± 32.01 | 124.26 ± 45.21 | 193.34 ± 44.85 | |
| Brain | 13.36 ± 2.35(a) | 18.89 ± 3.01(b) | 7.33 ± 2.97(c) | |
| Muscle | 3.58 ± 0.35(a) | 2.89 ± 0.21(ab) | 1.8 ± 0.13(b) | |
| AChE | Brain | 9.66 ± 0.72(a) | 6.51 ± 0.52(b) | 10.47 ± 0.44(a) |
| Muscle | 0.55 ± 0.21 | 1.93 ± 0.44 | 1.01 ± 0.11 | |
| GST | Gills | 0.04 ± 0.009 | 0.04 ± 0.003 | 0.23 ± 0.01 |
| Liver | 0.50 ± 0.25(a) | 0.90 ± 0.44(a) | 3.71 ± 0.18(b) | |
| AST | Liver | 10.47 ± 1.64 | 11.69 ± 0.59 | 12.75 ± 2.31 |
| ALT | Liver | 2.03 ± 0.71 | 1.97 ± 0.75 | 3.35 ± 0.53 |
| AST/ALT | Liver | 5.15 ± 2.26 | 5.94 ± 0.75 | 3.38 ± 2.76 |
| ALP | Liver | 17.78 ± 8.71 | 32.99 ± 13.09 | 18.7 ± 5.37 |
| Treatment | Control | T1 | T2 | |
|---|---|---|---|---|
| Biomarker | Organ | |||
| CAT | Gills | 5.94 ± 0.91(ab) | 2.56 ± 0.51(a) | 8.32 ± 1.92(b) |
| Liver | 74.74 ± 32.01 | 124.26 ± 45.21 | 193.34 ± 44.85 | |
| Brain | 13.36 ± 2.35(a) | 18.89 ± 3.01(b) | 7.33 ± 2.97(c) | |
| Muscle | 3.58 ± 0.35(a) | 2.89 ± 0.21(ab) | 1.8 ± 0.13(b) | |
| AChE | Brain | 9.66 ± 0.72(a) | 6.51 ± 0.52(b) | 10.47 ± 0.44(a) |
| Muscle | 0.55 ± 0.21 | 1.93 ± 0.44 | 1.01 ± 0.11 | |
| GST | Gills | 0.04 ± 0.009 | 0.04 ± 0.003 | 0.23 ± 0.01 |
| Liver | 0.50 ± 0.25(a) | 0.90 ± 0.44(a) | 3.71 ± 0.18(b) | |
| AST | Liver | 10.47 ± 1.64 | 11.69 ± 0.59 | 12.75 ± 2.31 |
| ALT | Liver | 2.03 ± 0.71 | 1.97 ± 0.75 | 3.35 ± 0.53 |
| AST/ALT | Liver | 5.15 ± 2.26 | 5.94 ± 0.75 | 3.38 ± 2.76 |
| ALP | Liver | 17.78 ± 8.71 | 32.99 ± 13.09 | 18.7 ± 5.37 |
Note. Superscript letters indicate statistical differences between groups. Groups sharing the same letter are not significantly different, while groups with different letters differ significantly (analysis of variance p < 0.05, Tukey’s test p < 0.05). The values are expressed as means ± SD. T1 = 0.15 mg/L; T2 = 1.5 mg/L; AChE = acetylcholinesterase; ALT = alanine aminotransferase; ALP = alkaline phosphatase; AST = aspartate aminotransferase; CAT = catalase; GST = glutathione S-Transferase.
Enzyme activities (nkat/mg prot) in different tissues of Cnesterodon decemmaculatus exposed to chlorantraniliprole.
| Treatment | Control | T1 | T2 | |
|---|---|---|---|---|
| Biomarker | Organ | |||
| CAT | Gills | 5.94 ± 0.91(ab) | 2.56 ± 0.51(a) | 8.32 ± 1.92(b) |
| Liver | 74.74 ± 32.01 | 124.26 ± 45.21 | 193.34 ± 44.85 | |
| Brain | 13.36 ± 2.35(a) | 18.89 ± 3.01(b) | 7.33 ± 2.97(c) | |
| Muscle | 3.58 ± 0.35(a) | 2.89 ± 0.21(ab) | 1.8 ± 0.13(b) | |
| AChE | Brain | 9.66 ± 0.72(a) | 6.51 ± 0.52(b) | 10.47 ± 0.44(a) |
| Muscle | 0.55 ± 0.21 | 1.93 ± 0.44 | 1.01 ± 0.11 | |
| GST | Gills | 0.04 ± 0.009 | 0.04 ± 0.003 | 0.23 ± 0.01 |
| Liver | 0.50 ± 0.25(a) | 0.90 ± 0.44(a) | 3.71 ± 0.18(b) | |
| AST | Liver | 10.47 ± 1.64 | 11.69 ± 0.59 | 12.75 ± 2.31 |
| ALT | Liver | 2.03 ± 0.71 | 1.97 ± 0.75 | 3.35 ± 0.53 |
| AST/ALT | Liver | 5.15 ± 2.26 | 5.94 ± 0.75 | 3.38 ± 2.76 |
| ALP | Liver | 17.78 ± 8.71 | 32.99 ± 13.09 | 18.7 ± 5.37 |
| Treatment | Control | T1 | T2 | |
|---|---|---|---|---|
| Biomarker | Organ | |||
| CAT | Gills | 5.94 ± 0.91(ab) | 2.56 ± 0.51(a) | 8.32 ± 1.92(b) |
| Liver | 74.74 ± 32.01 | 124.26 ± 45.21 | 193.34 ± 44.85 | |
| Brain | 13.36 ± 2.35(a) | 18.89 ± 3.01(b) | 7.33 ± 2.97(c) | |
| Muscle | 3.58 ± 0.35(a) | 2.89 ± 0.21(ab) | 1.8 ± 0.13(b) | |
| AChE | Brain | 9.66 ± 0.72(a) | 6.51 ± 0.52(b) | 10.47 ± 0.44(a) |
| Muscle | 0.55 ± 0.21 | 1.93 ± 0.44 | 1.01 ± 0.11 | |
| GST | Gills | 0.04 ± 0.009 | 0.04 ± 0.003 | 0.23 ± 0.01 |
| Liver | 0.50 ± 0.25(a) | 0.90 ± 0.44(a) | 3.71 ± 0.18(b) | |
| AST | Liver | 10.47 ± 1.64 | 11.69 ± 0.59 | 12.75 ± 2.31 |
| ALT | Liver | 2.03 ± 0.71 | 1.97 ± 0.75 | 3.35 ± 0.53 |
| AST/ALT | Liver | 5.15 ± 2.26 | 5.94 ± 0.75 | 3.38 ± 2.76 |
| ALP | Liver | 17.78 ± 8.71 | 32.99 ± 13.09 | 18.7 ± 5.37 |
Note. Superscript letters indicate statistical differences between groups. Groups sharing the same letter are not significantly different, while groups with different letters differ significantly (analysis of variance p < 0.05, Tukey’s test p < 0.05). The values are expressed as means ± SD. T1 = 0.15 mg/L; T2 = 1.5 mg/L; AChE = acetylcholinesterase; ALT = alanine aminotransferase; ALP = alkaline phosphatase; AST = aspartate aminotransferase; CAT = catalase; GST = glutathione S-Transferase.
In this context, AChE activity in the brain was significantly inhibited at T1 (6.51 ± 0.52 nkat/mg prot vs. C: 9.66 ± 0.72 nkat/mg prot; p < 0.05), whereas no significant differences were observed between T2 (10.47 ± 0.44 nkat/mg prot) and the control group. In contrast, AChE activity in muscle showed no significant changes across treatments (C: 0.55 ± 0.21; T1: 1.93 ± 0.44; T2: 1.01 ± 0.11 nkat/mg prot).
Catalase activity also exhibited tissue-specific patterns. In the gills, significant inhibition was observed at T1 (2.56 ± 0.51 nkat/mg prot vs. C: 5.94 ± 0.91 nkat/mg prot; p < 0.05), followed by an increase at T2 (8.32 ± 1.92 nkat/mg prot; p < 0.05 vs. T1), although T2 did not differ significantly from the control group. In the brain, CAT activity displayed a biphasic response: stimulation at T1 (18.89 ± 3.01 nkat/mg prot vs. C: 13.36 ± 2.35 nkat/mg prot; p < 0.05) and inhibition at T2 (7.33 ± 2.97 nkat/mg prot; p < 0.05 vs. C). On the other hand, muscle CAT activity decreased dose-dependently, with significant inhibition at T2 (1.8 ± 0.13 nkat/mg prot vs. C: 3.58 ± 0.35 nkat/mg prot; p < 0.05).
While CAT activity in the liver did not show significant changes among treatments (C: 74.74 ± 32.01; T1: 124.26 ± 45.21; T2: 193.34 ± 44.85 nkat/mg prot), GST activity increased significantly at T2 (3.71 ± 0.18 nkat/mg prot vs. C: 0.50 ± 0.25 and T1: 0.90 ± 0.44 nkat/mg prot; p < 0.05), showing a dose-dependent increase. In the gills, no significant differences in GST activity were detected among treatments (C: 0.04 ± 0.009; T1: 0.04 ± 0.003; T2: 0.23 ± 0.01).
Finally, no significant differences were observed in hepatic metabolic markers, including AST (10.47–12.75 nkat/mg prot), ALT (1.97–3.35 nkat/mg prot), AST to ALT ratio (3.38–5.94), or ALP (17.78–32.99 nkat/mg prot).
IBR index
The IBR values, integrating enzymatic and behavioral biomarker responses, revealed a dose-dependent stress gradient. The highest values were recorded in individuals exposed to T2, followed by T1, and the lowest in the control group.
For AChE, CAT, and GST, individuals exposed to T2 exhibited an IBR value of 6.16 (Figure 2A). Catalase activity showed the most significant alterations in the liver, gills, and muscles. In contrast, GST activity had a significantly high response in the gills, and AChE activity was primarily affected in the brain. These findings suggest that the highest CHL concentration induced a marked increase in CAT activity, with the gills being the most severely impacted organ. At T1 (IBR = 2.55), CAT activity in the brain and muscle and AChE activity in the brain showed values lower than those of the control group. In contrast, CAT activity in the brain, AChE activity in the muscle, and GST activity in the liver exhibited the highest stress responses among all treatments, highlighting a complex interplay of inhibitory and stimulatory effects at lower CHL concentrations.
Radar plots representing the Integrated Biomarker Response (IBR) for C: control group, T1: 0.15 mg/L and T2: 1.5 mg/L. (A) CAT = catalase. (CAT L): liver; (CAT B): brain; (CAT G): gill; (CAT M): muscle; AChE = acetylcholinesterase. (AChE B): brain; (AChE M): muscle; GST = glutathione S-transferase. (GST L): liver; (GST G): gill. (B). ALP = alkaline phosphatase; AST = aspartate aminotransferase; ALT = alanine aminotransferase. (C) As = average speed; Dt = distance traveled; Im = duration of immobility; Ms = maximum speed.
Regarding transaminases (Figure 2B), individuals exposed to T2 (IBR = 1.71) showed a significant hepatic impact, with increased AST and ALT activities. Conversely, in T1 (IBR = 0.82), the liver-specific enzyme ALP displayed the highest response.
For behavioral parameters (Figure 2C), the IBR index was highest in T2 (IBR = 6.33), followed by T1 (IBR = 3.89) and the control group (IBR = 0.03). Individuals exposed to 1.5 mg/L CHL exhibited the most pronounced behavioral effects, including the shortest distance traveled, the lowest average and maximum speed, and the most extended immobility episodes.
Discussion
Our results indicate that exposure to sublethal concentrations of CHL induces multiple effects in C. decemmaculatus across several levels of response levels. This study represents the first report on the toxicity of the CHL diamide insecticide in this key aquatic South American bioindicator species.
The most significant effects were observed in the locomotor activity of individuals, resulting in marked hypoactivity. Individuals exposed to the highest concentration of CHL traveled approximately one-third the distance of the control group. They exhibited average and maximum speeds of approximately half that of the control group. Furthermore, they displayed nearly a threefold increase in immobility compared with the control group. Among the enzymatic biomarkers analyzed, AChE, CAT, and GST exhibited different response patterns depending on the concentration of CHL and the organs analyzed, revealing a complex, dose-dependent relationship with both inhibitory and stimulatory effects at low concentrations.
Pesticides can impact fish locomotor activity through physiological and neurological mechanisms, including imbalances in energy metabolism, disruption of nerve signal transmission, muscle dysfunction, and alterations in sensory sensitivity (Scott & Sloman, 2004; Sharma et al., 2019). Behavioral impairments are commonly linked to changes in cholinesterase activity (Scott & Sloman, 2004). Acetylcholinesterase inhibition reduces the hydrolysis of acetylcholine, leading to excessive stimulation of cholinergic receptors, which can result in muscle spasms, weakness, paralysis, and decreased survival rates (Bernal-Rey et al., 2020; Slaninova et al., 2009).
In C. decemmaculatus, exposure to organophosphates, pyrethroids, and carbamates has been shown to reduce swimming velocity and distance traveled, associated with AChE inhibition (Bernal-Rey et al., 2020; Bonifacio et al., 2016, 2017; De la Torre et al., 2005). However, in our study, AChE activity in muscle tissue remained similar between the control group and the treatments, suggesting no significant anticholinergic effect in this tissue. In addition, AChE activity decreased in the brain only in individuals exposed to the lowest concentration.
These results suggest that locomotor activity alterations are more likely linked to muscle dysfunction caused by diamides’ RyR-mediated toxicity rather than to direct neurotoxic effects. While diamides exhibit greater selectivity for insect RyRs, our findings indicate that nontarget organisms, such as fish, may experience similar effects. However, for a direct assessment of the relationship between the observed locomotor alterations and the mode of action of diamides, it would be particularly useful to evaluate subcellular molecular pathways, such as RyR gene expression and genes related to cellular Ca2+ homeostasis. In this regard, Stinson et al. (2022) reported significant changes in the expression of genes associated with detoxification and neuromuscular function in Pimephales promelas exposed to 2.40 mg/L of CHL (a concentration close to T2); however, they did not assess its effects on locomotor activity, for which our study provides the first record.
The activity of detoxification and biotransformation enzymes also revealed physiological effects in exposed individuals, which may exacerbate the observed locomotor changes. In this context, CAT in muscle and brain tissue exhibited the highest susceptibility to oxidative stress, indicating a reduced capacity to neutralize hydrogen peroxide (H2O2), potentially leading to the accumulation of ROS (Santana et al., 2022). In muscle tissue, such oxidative damage could directly impair muscle fibers and reduce contraction efficiency, whereas, in the brain, CAT inhibition might disrupt neuronal redox balance, indirectly affecting locomotor activity. The hormetic response observed in the brain has been documented in fish exposed to insecticides (Amenyogbe et al., 2021; Bonifacio et al., 2017; Fulton & Key, 2001; Rodrigues et al., 2015); however, for chlorinated hydrocarbons (CHL), this phenomenon has only been reported in insects and anurans (Fonseca Peña & Brodeur, 2023; Tuelher et al., 2017; Wang et al., 2022). The stimulatory response observed at low CHL doses suggests an adaptive mechanism to mitigate oxidative stress induced by CHL exposure. At higher concentrations, however, CAT inhibition overwhelms this compensatory mechanism, leading to potential cellular damage.
The increase in GST activity suggests an adaptive response in the liver, a key organ in detoxification (Mohamed et al., 2022). However, no significant differences were found in CAT and transaminase activity between treatments and the control group, despite their known sensitivity as biomarkers of insecticide exposure in C. decemmaculatus (Bonifacio & Hued, 2019; Bonifacio et al., 2016; Ossana et al., 2019).
All these results indicate a complex interaction between exposure and the organisms’ physiological responses. In this context, the IBR index integrates the broad variability observed into a single value, providing a clearer understanding of the effects. It revealed a dose-dependent relationship, with individuals exposed to T2 exhibiting more than twice the IBR value of those exposed to T1 and nearly four times that of the control group.
Chlorantraniliprole exposure negatively impacts fish behavior and physiology. These results show that an insecticide with a specific mechanism of action for insects can have serious consequences on aquatic organisms, leading to reduced survival and growth rates with significant ecological consequences.
Current research on the impact of CHL on fish is limited, mainly focusing on a few species from Southeast Asia. Our work is the first report on South American native fishes, where this compound is widely used. Moreover, our findings reveal that even at concentrations lower than those typically observed in South American aquatic ecosystems, which range from trace detections to a maximum of 10.2 μg/L (Navarro et al., 2024; Rodríguez-Bolaña et al., 2023), CHL exposure induces significant adverse effects in nontarget organisms. Furthermore, the sublethal concentration approach employed in this study provided valuable insights into the toxicity mechanisms of diamides and the efficacy of using multiple biomarkers to assess biological responses in exposed organisms. We encourage future studies to include chronic toxicity tests to determine the impact of CHL at both individual and ecological levels. We also recommend conducting studies with environmentally relevant concentrations and evaluating toxicological interactions with other insecticides, such as organophosphates, which frequently co-occur in South American aquatic environments. These studies can provide crucial information on toxicity under more realistic scenarios and may be critical for establishing CHL regulatory limits and environmental management measures.
Conclusion
Our study concludes that sublethal acute exposure to CHL is toxic to C. decemmaculatus, as evidenced by alterations in locomotor activity and dose-dependent responses in antioxidant and detoxification enzyme activities. Adaptive responses involving AChE, CAT, and GST were observed, highlighting the organism’s ability to manage oxidative stress. The most evident effect was the reduction in locomotor activity, which appears to be more closely related to CHL’s mode of action than cholinergic effects. These demonstrate that insect-specific compounds like diamides can severely affect nontarget species. Such impairments may lead to diminished endurance, slower swimming speeds, and increased resting periods, potentially impacting feeding efficiency, reproductive success, predator avoidance, and survival rates. Behavioral parameters appear to be more sensitive biomarkers than conventional AChE activity for assessing the effects of muscle-targeting insecticides in fish.
Future studies should incorporate the analysis of molecular biomarkers, such as RyR gene expression, for a more direct evaluation of these effects. Chlorantraniliprole is the most widely used diamide insecticide globally and has a high potential for runoff into aquatic ecosystems; these findings highlight the need for regular monitoring programs to assess and mitigate its risks in aquatic environments.
Data availability
The data supporting the findings of this study are available from the corresponding author upon request.
Author contributions
César Rodríguez-Bolaña (Conceptualization, Data curation, Formal analysis, Investigation, Methodology, Visualization, Writing—original draft, Writing—review & editing), Andrés Pérez-Parada (Conceptualization, Funding acquisition, Investigation, Supervision, Writing—review & editing), Andrea Hued (Methodology, Resources, Supervision, Writing—review & editing), Alejo Bonifacio (Data curation, Formal analysis, Methodology, Writing—review & editing), Marina Tagliaferro (Data curation, Formal analysis, Methodology, Writing—review & editing), and Franco Teixeira de Mello (Conceptualization, Funding acquisition, Investigation, Supervision, Writing—review & editing)
Funding
C.R.-B. received support from Becas de Finalización de Posgrado CAP (Comisión Académica de Posgrado). C.R.-B., A.P.-P., and F.T.d.M. received support from the SNI (Agencia Nacional de Investigación e Innovación, ANII, Uruguay).
Conflicts of interest
None declared.
Acknowledgments
C.R.-B., A.P.-P., and F.T.d.M. thank Sistema Nacional de Investigadores (SNI) and the Programa de Doctorado en Ciencias Ambientales (Udelar).
Ethical statement
This study was conducted in compliance with national and institutional guidelines for animal research. No specific approval was required.
