Toxicity of Commonly Used Plasticizers to the Freshwater Organisms Tilapia sparrmanii (Fish) and Caridina nilotica (Shrimp): Lethal and Sublethal Effects

ABSTRACT

The toxicity of microplastics in aquatic environments is usually due to plasticizers, the chemical additives that keep the plastic polymers together. Thus, the current study reports on the toxicity of three common plasticizers found in freshwater ecosystems and their impacts on two South African freshwater organisms at the organismal and biochemical levels. Tilapia sparrmanii (fish) and Caridina nilotica (shrimp) were exposed to varying concentrations of the test plasticizers, including bisphenol‐A (BPA), calcium stearate (CAS), and dibutyl phthalate (DBP). The impacts of these plasticizers on mortality and biomarkers (acetylcholinesterase activity and lipid peroxidation) were investigated using 96‐h short‐term static nonrenewal and 21‐day long‐term static renewal exposure methods, respectively. All experiments were conducted in temperature‐controlled rooms. Mortality was determined after 96 h, while biochemical effects were measured after 21 days. The results revealed that all three plasticizers significantly affected the mortality of both organisms. Also, acetylcholinesterase activity per unit protein in shrimp decreased significantly at all levels of exposure, while lipid peroxidation increased significantly at all levels of exposure. This study has shown that short‐term and long‐term exposures to the tested plasticizers could adversely impact populations of the tested organisms at both the organismal and biochemical levels.

1. Introduction

The ubiquity of plastics of all sizes in aquatic environments has attracted global concern about their ecological and human health effects [1, 2, 3]. Over the past few years, plastic production has grown significantly [4, 5, 6], and about 60% of these plastics end up in the environment where they accumulate and become physical and chemical stressors [7, 8]. As chemical stressors, they may act as vectors for absorbed chemicals and leach plasticizers (additives in plastics that increase their efficiency by making them more flexible and tougher). Leaching of plasticizers from plastic polymers has commonly been noted in a process that introduces the plasticizer to the environment around the plastic and to embrittlement of the plastic [8, 9, 10]. Plasticizers can leach when they have not formed chemical bonds with the polymer with which they are dispersed [9, 11, 12]. The introduction of plasticizers into the environment may be a major environmental threat that can affect both humans and organisms [13]. Calcium stearate (CAS), phthalate, and bisphenol‐A (BPA) are common plasticizers in polypropylene [14], polyvinyl chloride [15], and polyethylene [16], respectively, that have been reported in aquatic environments [17, 18, 19]. These plasticizers are added to plastics to make them more flexible and durable; these plastics generally contain up to 50% plasticizers by weight [20]. Phthalates are also additives in fragrances, paints, sealants, cardboard, adhesives, lubricants, inks, and pesticides [21]. However, phthalates are not physically bound to the plastic polymer and can easily leach into the environment during manufacture, use, and disposal [22, 23]. Plasticizers such as phthalate and BPA have been reported widely in various environmental matrices [24]. Dibutyl phthalate (DBP) has been identified in the European Union as a substance of very high concern because of its ubiquity in the environment and toxic properties [19].

Plasticizers in the environment can affect organisms and humans through various pathways such as ingestion, inhalation, and skin absorption/contact [25, 26, 27, 28]. Most plasticizers are regarded as endocrine‐disruptive chemicals (EDCs) and can cause substantial harm to the respiratory, reproductive, and endocrine systems of mammals [29, 30, 31]. In the freshwater environment, plasticizers can leach and persist at high concentrations [32, 33], causing severe effects on the ecosystem and human health via the food chain. Effects associated with these compounds, such as reduced fertility, feminization, reproductive organ abnormalities, or altered sexual behavior, have not been observed only in mammals but also in fishes and other aquatic organisms [34].

Due to the continuously high detection rates of plasticizers such as DBP and BPA in the aquatic environment, many studies have focused on the toxicity of these substances. However, the major challenge concerning these studies is that they are often conducted at environmentally unrealistic concentrations, which in turn makes such findings not relevant ecologically [35]. Further, [35] argued for long‐term ecotoxicological experiments with environmentally realistic concentrations to be assessed through ecologically relevant endpoints.

The freshwater shrimp Caridina nilotica (Decapoda: Atyidae) is a widely distributed species found in the inland waters of South Africa. It is used as a standard organism for toxicity testing in South Africa, providing ecologically relevant data on environmental contaminants [36, 37]. Tilapia sparrmanii (Banded Tilapia) is a widely distributed and adaptable cichlid species found in warmer freshwater habitats throughout South Africa. It exhibits sensitivity to pollutants, making it a valuable bioindicator species. The physiological responses of this fish to toxic substances can yield critical biomarkers that provide early warning signs of ecosystem degradation. These indicators are essential for assessing pollution‐associated ecological risks, highlighting the importance of monitoring this species to protect freshwater environments [38].

Plasticizers'pollution of the aquatic ecosystem may interfere with ecosystem health and pose numerous risks, including neurotoxicity and oxidative stress to aquatic life [39]. Acetylcholinesterase (AChE) and lipid peroxidation (LPx) are necessary biomarkers for evaluating the health status of aquatic organisms and understanding the broader ecological implications of chemical exposure. Therefore, in this study, AChE and LPx activities were used as endpoint biomarkers to provide valuable insights into neurotoxic effects and oxidative stress induced by plasticizers, respectively. To fill the research gap of paucity of studies on the effects of plasticizers on South African freshwater ecosystems, this study investigated the ecotoxicological impacts of CAS, DBP, and BPA on the South African riverine organisms C. nilotica and T. sparrmanii after exposure to environmentally relevant concentrations. The specific objectives were to: (1) determine the lethal concentrations (LCx) of the tested plasticizers on T. sparrmanii and C. nilotica over 96‐ and 21‐day exposure periods; (2) investigate sublethal impacts of plasticizer exposure on biomarkers in both species to understand potential long‐term ecological consequences; (3) establish dose–response curves for both lethal and sublethal effects to explain how varying concentrations of plasticizers affect the health and survival of these organisms.

2. Materials and Methods

2.1. Test Chemicals

The plasticizers tested in this study were BPA ((CH₃)₂C(C₆H₄OH)₂, molecular weight: 228.29), CAS ([CH₃(CH₂)₁₆COO]₂Ca, molecular weight: 607.02), and DBP (C₆H₄‐1,2‐[CO₂(CH₂)₃CH₃]₂, molecular weight: 278.34), with Chemical Abstracts Service Registry Numbers (CAS RN) 80‐05‐7, 1592‐23‐0, and 84‐74‐2, respectively. Analytical grades BPA, CAS, and DBP all with respective analytical assays of 97% or more, were purchased from commercial sources (Sigma‐Aldrich, USA). Test concentrations of the plasticizers BPA and DBP were obtained from a literature review of reported environmental concentrations [35, 40, 41, 42] and laboratory‐based range‐finding tests. Range‐finding tests were used for CAS to determine exposure concentrations due to unavailable information on freshwater environments. Dechlorinated tap water, passed through activated charcoal, was used as a diluent and control treatment for all three plasticizers. However, DBP and CAS were first dissolved in acetone and ethylenediaminetetraacetic acid disodium salt dihydrate (EDTA), respectively, as they sparingly dissolved in water. Analytical grades of acetone (99.5% assay) and EDTA (99.0% assay) were purchased from commercial sources (Sigma‐Aldrich, USA). All reported concentrations in this study were nominal.

2.2. Test Organisms

The test organisms used in this study were juveniles of T. sparrmanii and C. nilotica , with mean lengths of 14.67 ± 1.75 mm and 8.73 ± 1.07 mm, respectively. Juveniles of the test organisms were used because early life stages are vital for understanding the full scope of ecological impacts, ensuring the protection of vulnerable populations within ecosystems. Both test organisms were obtained from cultures maintained in the wet lab of the Institute for Water Research, Rhodes University, in a 12‐h light:12‐h dark artificial light regime using Biolux fluorescent tubes in a temperature‐controlled room at 25°C (±0.05°C).

2.3. Exposure Experiments

All test organisms were acclimatized under laboratory test conditions for 48 h before exposure experiments began. The experimental protocol follows the various tests presented as part of DEEEP methods (i.e., Methods for Direct Estimation of Ecological Effect Potential) [43], whereby a 96‐h static nonrenewal test and a 21‐day static renewal test were employed for short‐term and long‐term exposures, respectively. For the short‐term 96‐h exposure tests, 10 individuals of T. sparrmanii and C. nilotica were separately exposed to different concentrations: BPA (0, 0.00025, 0.0005, 0.001, 0.002, 0.004 mg/L), CAS (EDTA control, 0, 100, 200, 300, 400, 500 mg/L), and DBP (acetone control, 0, 0.00014, 0.00027, 0.00054, 0.00108, 0.00215 mg/L). The experimental vessels used were 600‐mL glass beakers filled with different concentrations of the test chemical in triplicates. Test organisms were not fed during the experimental period, and dead organisms were recorded twice daily and removed from experimental vessels. Each experiment ended after 96 h of exposure. For the 21‐day sublethal test, 10 individuals each of T. sparrmanii and C. nilotica were separately exposed to different concentrations: BPA (0, 0.00025, 0.0005, 0.001 mg/L), CAS (0, 1.25, 2.5, 5 mg/L), and DBP (0, 0.00014, 0.00027, 0.00054 mg/L). These concentrations were relatively lower than those used for the short‐term exposure tests, so sublethal effects could be elicited and measured. Also, the solvent control was not added in the long‐term experiment as they did not record any mortality in the short‐term tests. The experimental vessels used were 600‐mL glass beakers filled with different concentrations of the test chemical in triplicates. Test solutions were changed every other day, and organisms were fed TetraMin tropical flakes daily and just after the change of solution. At the end of the 21‐day test period, all surviving shrimps were quickly transferred individually into Eppendorf tubes, immediately frozen in liquid nitrogen, and stored at −80°C in a refrigerator until biochemical analyses were measured.

The test concentrations used in both short‐term and long‐term studies were nominal, and experiments were conducted under similar conditions as the culture, that is, a 12‐h light:12‐h dark artificial light regime using Biolux fluorescent tubes in a temperature‐controlled room at 25°C (±0.05°C). The experimental chambers were provided with aeration, while water quality parameters, including pH, dissolved oxygen (DO), electrical conductivity (EC), and temperature, were measured in all concentrations at the beginning of experiments and after the change of solutions. The Rhodes University Animal Research Ethics Committee (RU‐AREC), which approves all animal experiments, approved the study, with approval number 2021‐0684‐6375.

2.4. AChE Assay

ACHEs are enzymes that hydrolyze the neurotransmitter acetylcholine to acetate and choline. Changes in AChE activity may result from exposure to chemical stressors, including certain insecticides. AChE activity was measured using an Assay Kit (Catalog Number MAK119, Sigma‐Aldrich, USA). This assay is an optimized version of the Ellman method in which thiocholine, produced by AChE, reacts with 5,5‐dithiobis(2‐nitrobenzoic acid) to form a colorimetric (412 nm) product, proportional to the AChE activity present. One unit of AChE is the enzyme that catalyzes the production of 1.0 millimole of thiocholine per minute at pH 7.5 at room temperature. Samples of T. sparrmanii or C. nilotica were separately defrosted to room temperature, then wholly macerated and homogenized in 0.1 M phosphate buffer at pH 7.5 to extract the proteins present. The extracts were clarified by centrifuging at 14 000 rpm for 5 min; 100 μL of supernatant was collected for the AChE assay and analyzed colorimetrically following kit instructions.

2.5. Lipid Peroxidase Assay

LPx is the degradation of lipids that occurs as a result of oxidative damage and is a useful marker for oxidative stress. Polyunsaturated lipids are susceptible to an oxidative attack, typically by reactive oxygen species (ROS), resulting in a well‐defined chain reaction with the production of end products such as malondialdehyde (MDA). Lipid peroxidase activity was assessed using an Assay Kit (Catalog Number MAK085, Sigma‐Aldrich, USA). In this kit, LPx is determined by the reaction of MDA with thiobarbituric acid to form a colorimetric product, proportional to the MDA present. Samples of T. sparrmanii or C. nilotica were separately defrosted to room temperature, then 10 mg macerated and homogenized on ice in 300 μL of MDA lysis buffer containing 3 μL of butylated hydroxytoluene. Samples were then centrifuged at 13 000g for 10 min; then 200 μL of supernatant were taken for further colorimetric analysis of lipid peroxidase following the kit's instructions.

2.6. Total Protein Assay

Levels of total protein in sampled material were also determined so that enzyme activity could be expressed in terms of enzyme activity per unit of protein present. The protein assay used an Assay Kit (Catalog Number 1.10306.0500, Merck, Germany). This used the Bradford method for assessing protein and is based on the property of Coomassie Brilliant Blue G‐250 to bind to proteins. When this happens, its absorbance maximum is shifted from 465 to 595 nm, and this provides a means of assessing protein levels. Samples of T. sparrmanii or C. nilotica were separately defrosted to room temperature, then wholly macerated and homogenized in 0.1 M phosphate buffer at pH 7.5 to extract the proteins present. The extracts were clarified by centrifuging at 14 000 rpm for 5 min; 100 μL of supernatant was collected for the protein assay.

2.7. Data Analysis

All mortality and biochemical data were tested for normality using the Kolmogorov–Smirnov test [44]. One‐way analysis of variance (ANOVA) was performed to test the hypothesis that all mean mortality and biochemical data values across different concentrations were equal. In the presence of significant observations, Tukey HSD multiple comparisons post hoc tests were used to compare the specific difference between any two means. Where data were counted and followed a Poisson distribution, comparisons used a general linear model process. A Chi‐squared test was used to compare the generated model with a null model so that the model's significance could be evaluated. Plasticizers (BPA, CAS, and DBP) 96‐h LCx (LC₅, LC₅₀, and LC₉₅) values and the associated 95% confidence limits were calculated for T. sparrmanii and C. nilotica using the ED function in the dose–response curve (drc) package in R. Pearson's correlation was used to compare the relationship between protein content and AChE or LPx activity. A Pearson's correlation value (R ²) of greater than 0.5 implies a strong and positive relationship. Statistical analyses were performed using R version 4.3.3 [45] together with the libraries ecotox [46], drc [47], multcomp [48], and ggplot2 [49]. All statistical decisions were made at alpha = 0.05 a priori.

3. Results

Mean water quality parameters with standard deviations (±SD) of T. sparrmanii and C. nilotica exposed to BPA, CAS, and DBP in short‐term and long‐term tests are presented in Table 1. These ranges of values were all within the acclimated conditions of the culture maintained in the laboratory.

TABLE 1.

Mean water quality parameters with standard deviations (±SD) of T. sparrmanii and C. nilotica exposed to bisphenol‐A, calcium stearate, and dibutyl phthalate in short‐term and long‐term tests.

Plasticizer	pH	DO (mg/L)	EC (mS/cm)	Temperature (°C)
Bisphenol‐A	6.25 ± 0.05	6.52 ± 0.86	0.54 ± 0.01	23.75 ± 0.44
Calcium stearate	8.14 ± 0.25	5.89 ± 0.61	0.72 ± 0.21	24.83 ± 0.27
Dibutyl phthalate	7.41 ± 0.16	5.89 ± 0.10	0.43 ± 0.07	24.50 ± 0.41

3.1. Short‐Term 96‐h Lethal Test

There were no mortalities in organisms exposed to the negative control, and there were no significant differences in the mortalities of organisms exposed to solvent and negative controls (p > 0.05). Thus, solvent control was removed from subsequent analysis. There were no mortalities in the control groups for the two exposed organisms (i.e., T. sparrmanii and C. nilotica ), but there were significant differences between control and treatment groups (p < 0.05). For T. sparrmanii , exposure to BPA caused significant mortalities in concentrations of 2 and 4 mg/L compared to all other groups, while mortalities in concentrations of 0.25, 0.5, and 1 mg/L were not significant from each other and control. There were no significant differences in the mortalities of T. sparrmanii exposed to 100–300 mg/L CAS compared to the control group, but mortalities in 500 mg/L were significantly higher than all other groups except mortalities in 400 mg/L. Similarly, there were significant differences in the mortality of T. sparrmanii exposed to 1–4 mg/L DBP compared to the control group, while mortalities in 0.25 and 0.5 mg/L were not significantly different from each other (Figure 1A). The associated dose–response curves are shown in Figure 2. For C. nilotica , exposure to BPA caused significant mortalities in 1 and 4 mg/L, but mortalities in 2 mg/L were significant from either concentration compared to control, while mortalities in 0.25 and 0.5 mg/L were not significantly different from the control group. Mortalities in all the CAS‐tested concentrations were significantly different from the control group, although there were no significant differences in the mortalities between 1 and 2 mg/L. Similarly, there were significant differences in mortality between all the DBP‐tested concentrations, although there were no significant differences between 100 and 200 mg/L compared to the control group (Figure 1B). The associated dose–response curves are shown in Figure 3. The mortalities at 96 h were used to estimate LCx and their 95% confidence limits for all three plasticizers.

FIGURE 1.

Mortality of T. sparrmanii (A) and C. nilotica (B) after 96 h exposed to varying concentrations of the tested plasticizers (BPA: 0–0.004 mg/L; CAS: 0–500 mg/L; DBP: 0–0.00215 mg/L).

FIGURE 2.

Dose–response curves ±95% confidence intervals (shaded regions) describing the changes in T. sparrmanii mortality exposed to bisphenol‐A (A), calcium stearate (B), and dibutyl phthalate (C).

FIGURE 3.

Dose–response curves ±95% confidence intervals (shaded regions) describing the changes in C. nilotica mortality exposed to bisphenol‐A (A), calcium stearate (B), and dibutyl phthalate (C).

The estimated LC₅, LC₅₀, and LC₉₅ values with 95% confidence limits for BPA, CAS, and DBP based on mortality after a 96‐h exposure are presented in Table 2 for T. sparrmanii and Table 3 for C. nilotica . At the level of LC₅, CAS and DBP seem to be more toxic to C. nilotica than to T. sparrmanii , while BPA seems to be more toxic to T. sparrmanii than to C. nilotica . However, at the median concentration level (LC₅₀), all three tested plasticizers proved to be more toxic to C. nilotica than to T. sparrmanii . Similar observations were made for the toxicity of these plasticizers at the LC₉₅ level.

TABLE 2.

Estimated LC₅, LC₅₀, and LC₉₅ values (in mg/L) with 95% confidence limits (presented in parentheses) for bisphenol‐A, calcium stearate, and dibutyl phthalate for T. sparrmanii based on mortality after a 96‐h exposure.

Plasticizer	LC5	LC50	LC95
Bisphenol‐A	0.0003 (0.0000–0.0006)	0.0067 (0.0035–0.0415)	0.1550 (0.0298–25.700)
Calcium stearate	405 (27.7–4.77e2)	726 (559–1.49e8)	1301 (763–6.87e14)
Dibutyl phthalate	0.0001 (0.0000–0.0002)	0.0005 (0.0004–0.0006)	0.002 (0.0014–0.0035)

TABLE 3.

Estimated LC₅, LC₅₀, and LC₉₅ values (in mg/L) with 95% confidence limits (presented in parentheses) for bisphenol‐A, calcium stearate, and dibutyl phthalate for C. nilotica based on mortality after a 96‐h exposure.

Plasticizer	LC5	LC50	LC95
Bisphenol‐A	0.0004 (0.0001–0.0007)	0.0017 (0.0012–0.0026)	0.0074 (0.0042–0.0283)
Calcium stearate	115 (58.6–155)	251 (200–300)	548 (429–920)
Dibutyl phthalate	0.0001 (0.0000–0.0002)	0.0003 (0.0002–0.0004)	0.0019 (0.0012–0.0040)

3.2. Long‐Term 21‐Day Biochemical Assay

3.2.1. AChE and Lipid Peroxidase Activities in T. sparrmanii

The dose–response curve for AChE activity for T. sparrmanii exposed to BPA, CAS, and DBP for 21 days is shown in Figure 4A. The figure shows increasing levels of plasticizers resulting in decreased AChE activities. Results of BPA and CAS have relatively low variation, which led to a significant model fit to the data (p < 0.05). The response of AChE to DBP in fish also showed decreasing levels of the plasticizer and relatively little variation in the data. Despite this, the statistical model's fit is insignificant (p > 0.05). In T. sparrmanii , statistically significant differences between plasticizer exposures in the expression of AChE were found. The enzyme responded differently to the different plasticizers (p < 0.05), and also to differing concentrations of the various plasticizers (p < 0.05). Not surprisingly, strong statistical support for the derived model was found (p < 0.05).

FIGURE 4.

Dose–response curves of acetylcholinesterase activity (A) and lipid peroxidase activity (B) in T. sparrmanii exposed to varying concentrations of the tested plasticizers (BPA: 0–0.001 mg/L; CAS: 0–5 mg/L; DBP: 0–0.00054 mg/L).

The response of lipid peroxidase in T. sparrmanii to exposure to BPA, CAS, and DBP for 21 days is shown in Figure 4B. The reaction of lipid peroxidase activity in T. sparrmanii to BPA increased as concentration levels of the plasticizer also increased. Similarly, CAS and DBP increased lipid peroxidase levels in T. sparrmanii exposed to increasing levels of these plasticizers. For each of these two plasticizers, the model fit of the curve was statistically significant (p < 0.05).

3.2.2. AChE and Lipid Peroxidase Activities in C. nilotica

The dose–response curves of AChE activity in juvenile C. nilotica exposed to CAS, BPA, and DBP for 21 days are shown in Figure 5A. For BPA, the fitted curve, which was used to estimate effective concentration (EC_x) values, exhibited a roughly linear relation between log concentration and AChE activity. It was clear that a greater response of AChE activity was found at 0.25 mg/L than at 1.0 mg/L, which adds to the variation in the response. The overall AChE activity was high at 4 mg/L BPA, and it seems that this enzyme activity is stimulated overall by BPA. The line fit was not statistically significant (p > 0.05). For CAS, the fitted curve reflected an apparent log‐logistic dose–response, and it is used to estimate the EC_x. The goodness of fit test for this regression model showed it to be not a significant fit (p > 0.05). The dose–response curve of AChE activity in juvenile C. nilotica exposed to DBP for 21 days revealed a slow increase in AChE activity with increasing DBP, with one data point at 0.00215 mg/L concentration of DBP having notably more AChE activity than the other. The fitted curve, which was significantly influenced by the high 0.00215 mg/L response, shows a roughly log‐logistic response to DBP. This model, which was the best of several assessed, provided estimates for EC_x.

FIGURE 5.

Dose–response curves of acetylcholinesterase activity (A) and lipid peroxidase activity (B) in C. nilotica exposed to varying concentrations of the tested plasticizers (BPA: 0–0.001 mg/L; CAS: 0–5 mg/L; DBP: 0–0.00054 mg/L).

The dose–response curves of lipid peroxidase activity in juvenile C. nilotica exposed to calcium BPA, CAS, and DBP for 21 days are shown in Figure 5B. The dose–response curve of lipid peroxidase activity in juvenile C. nilotica exposed to BPA exhibited the usual curve in that increased BPA concentrations seem to increase lipid peroxidase activity. The data was fitted to a log‐logistic model and were not a statistically significant fit (p > 0.05). The fitted curve for CAS showed an exponential increase in lipid peroxidase activity with increased exposure to CAS. Compared to the controls, no significant change in lipid peroxidase activity occurred at 1.25 mg/L of CAS, but after that, activity increased. The fitted model that was used to generate EC_x value estimates exhibited a statistically insignificant fit (p > 0.05). The dose–response curve of lipid peroxidase activity in juvenile C. nilotica exposed to DBP showed a trend apparent in the responses of the enzyme systems assessed to increasing levels of plasticizer in that lower levels had little effect on the enzyme system, but strong responses were observed at higher levels. As a result of the strong response, the log‐logistic dose–response curve fitted best and was used to generate the EC_x values.

4. Discussion

4.1. Short‐Term 96‐h Lethal Test

Under natural conditions, all the tested plasticizers can undergo biodegradation in water with a mean half‐life of 2.5–4 days for BPA [16, 50] and 1 day for DBP [15, 18]. If biodegradation proceeded at the same rate in laboratory cultures in dechlorinated water, the concentrations at the test's end would significantly differ. Thus, the change of solution at the rate of any other day ensured that the concentrations used in the experiments were maintained to a large extent. Also, no signs of recovery in test taxa were noted over time, suggesting that the tested plasticizer's degradation was limited throughout the experiments.

All the plasticizers assessed elicited a toxic response in test organisms at environmentally relevant levels during a relatively short 96‐h exposure. The results of exposure to BPA, CAS, and DBP varied between the test organisms but were not greatly different. For C. nilotica , DBP was the most toxic, followed by BPA and then CAS, which was far less harmful. Both CAS and DBP are largely insoluble, requiring treatment to solubilize so that the test could be completed. This treatment increased the amount of plasticizer in solution to an extent that it would be above the amount that could be normally present in water without some form of solubilizing treatment. In toxicity testing using a short exposure, this was desirable so that short‐term toxicity could be assessed. Without the treatment, both plasticizers would not have returned toxicity endpoints. This is particularly the case with a test using mortality as an endpoint, rather than a potentially more sensitive endpoint like a biochemical or reproductive endpoint. Little data were available on CAS toxicity [14, 18], but results from assessments of DBP toxicity to freshwater taxa [51, 52, 53] generally used endpoints more sensitive than mortality and returned lower effect concentrations than the ones produced here.

The derived mortality endpoints from this study indicate the plasticizers' direct, short‐term effect on mortality in the tested freshwater organisms. However, as indicated above, this required a means of solubilizing the plasticizers to levels greater than their normal solubility in water [14, 15]. The result was that mortality occurred in the test exposures; however, in the case of CAS, they were at levels that would not naturally occur owing to lack of solubility. This means that for CAS, none of the tested organisms would experience CAS‐mediated mortality in the environment.

DBP also elicited a toxic response in the tested organisms. As the effect concentrations were below the solubility limit, these might occur without being limited by the solubility of the plasticizer. However, environmental levels reported by [32] from surface freshwater range around 1 μg/L, with no levels exceeding 10 μg/L. This is below the majority of NOECs or PNECs reported in [53] and below any of the effect concentrations reported in this study. DBP has a biodegradation half‐life of 1 day in natural water, so it is unlikely to accumulate to significantly higher levels [53]. DBP was found in higher levels in water from polyethylene terephthalate water bottles [54, 55], and this may have consequences for water in contact with polymers containing DBP.

The lack of water solubility complicated toxicological analyses of CAS and DBP. Both plasticizers needed treatments to dissolve the plasticizer in water, which required the addition of EDTA or acetone to the test solution. The use of controls with EDTA and acetone assessed the impact of these components and found no detectable effect at the concentrations used. However, in trials, a higher level of EDTA had a toxic effect, and it is possible that some stress was caused by EDTA. Nevertheless, this remains speculative, and no behavioral changes were recorded at the dose used in this assay. The more concentrated CAS solutions were cloudy, which may have been a stressor in its own right, the more complicated some of the observations. DBP solutions were not cloudy, and no change in mortality or behavior in acetone controls was found. However, acetone is moderately toxic in its own right [56], and it may have stressed the test taxa to some extent.

The LC₅, LC₅₀, and LC₉₅ estimates for exposure to CAS indicate that at high enough concentrations, CAS can be lethal to this organism. However, it is important to note that to solubilize CAS, EDTA was added to the test solution. The use of EDTA raised the solubilized level of CAS above its normal solubility in water (i.e., 40 mg/L at 15°C; [14]), and as a result, the data represent levels of CAS that would not naturally be encountered in freshwater. Even the LC₅ value, which is often used in generating protective guidelines, as it is a level that would protect 95% of the exposed population, is greater for both taxa than the natural solubility level of CAS. C. nilotica showed a slightly greater toxic response compared with T. sparrmanii . CAS did not show greater toxicity compared with the other plasticizers even at higher tested levels. This lack of toxicity is, perhaps, not surprising given that CAS is the calcium salt of a naturally occurring fatty acid.

DBP was most toxic to C. nilotica and less toxic to T. sparrmanii . However, compared with CAS, DBP was considerably more toxic than CAS to both organisms. These results confirm the toxicity of DBP for freshwater organisms reported in [15] but also illustrate that responses to DBP by freshwater taxa may vary widely. The results from this assay concord with those from CAS in that DBP is only slightly water‐soluble [15]. However, the effect concentrations presented in Tables 1 and 2 for DBP show responses for concentrations lower than the limit set by the solubility level in water, and therefore, concentrations that are possible in the environment might elicit a mortality response.

4.2. Long‐Term 21‐Day Biochemical Assay

Plasticizer exposure can induce toxicity in aquatic organisms through multiple interconnected mechanisms. One primary pathway involves the inhibition of AChE, a crucial enzyme responsible for the breakdown of acetylcholine at synaptic junctions. Inhibition of AChE leads to the accumulation of acetylcholine, resulting in prolonged neural excitation, impaired behavior, muscle paralysis, and potentially fatal outcomes [57, 58]. Concurrently, plasticizers such as CAS, DBP, and BPA are known to stimulate the overproduction of ROS, which overwhelm cellular antioxidant defenses and lead to oxidative stress [59]. A key consequence of oxidative stress is LPx, where ROS attack polyunsaturated fatty acids in cell membranes, causing structural damage, loss of membrane integrity, and ultimately cell death [60]. Furthermore, oxidative stress can exacerbate neurotoxicity by further impairing AChE function, establishing a feedback loop that amplifies organismal damage [61]. Together, these processes explain the observed sublethal and lethal impacts of plasticizer exposure on aquatic species such as T. sparrmanii and C. nilotica .

BPA levels have been reported from the environment at levels around the higher levels assessed in this study and at higher levels in industrial effluents and recycling leachate in China, Japan, Germany, and Canada [62]. Significant levels of other bisphenol analogs have also been reported [63]. Bisphenol analogs (BPs) are increasingly used in various industries and are frequently detected in surface water, sediment, sewage, and sludge. The toxicity of bisphenol analogs, including BPA, has been reported in sewage effluents and is often higher in sludge and sediment than in water. Exposure to these compounds may disrupt endocrine functions, affecting thyroid hormones and potentially leading to cellular dysfunction and genetic damage [64, 65]. In another study, the risks posed by bisphenol compounds were evaluated by [66]. They reported that the risk quotient revealed a low to medium risk to algae, daphnia, and fish in irrigation rivers, whereas over 90% of soil sampling sites faced medium or high risk, with BPA being a major contributor to the risk. In another study, [67] evaluated the distribution, composition, and risk assessment of 8 endocrine‐disrupting chemicals (EDCs), including BPA in the surface waters. They also reported that fish were the most sensitive aquatic organisms and that BPA was among the compounds that caused higher risk. All these studies agree with [68] who also reported significant bisphenol compounds in surface water and sediment samples of a lake ecosystem, as well as in the muscle and gill tissues of fish. At the biochemical level, [69] opined that exposure of freshwater crayfish to bisphenol compounds resulted in oxidative stress by inducing elevated levels of ROS and inhibiting the activity of antioxidant‐related enzymes. They further asserted that bisphenol compounds could cause increased lipid content in the serum and hepatopancreas, which was associated with elevated lipid‐related enzyme activity and increased expression of related genes. Also, there were decreased levels of phosphatidylcholine (PC) and phosphatidylinositol (PI), disrupted glycerophospholipid (GPI) metabolism, and lipid deposition in the hepatopancreas due to the exposure. This indicates that the responses to BPA presented in this study are likely in polluted freshwater sites, though not at other locations. If BPA is derived from microplastics, the impact on river biota would depend on the plasticizer's high enough leach rate from the microplastic. BPA has been found to have endocrine effects [70, 71], and DBP may have endocrine activity [72]. The tests undertaken here assessed the effect of the plasticizers on enzyme systems that control acetylcholine management and lipid oxidation and so do not reflect directly on endocrine action. However, it may contribute to changes in the enzyme systems assessed.

Background information on CAS was not easily found. This plasticizer has many uses, including as an ingredient in food and pharmaceutical products, as a lubricant, use in waterproofing, and production of pulp and paper [11, 73, 74]. The United States Environmental Protection Agency (USEPA) classifies CAS as a chemical for sealants, anti‐adhesive agents, fillers, finishing agents, flame retardants, hydrophobic agents, intermediates, lubricant additives, neutralizing agents, polymer stabilizers, processing aids, and surface‐active agents [14]. CAS pellets in wastewater treatment work effluent were reported by [75, 76], but these were not quantified, and the amount of dissolved CAS is unknown. Although the use of CAS in food [14] suggests limited or no toxicity, the results of this study suggest that there may be an adverse effect on freshwater organisms.

Phthalates are among the major plasticizers that are widely spread and readily released into the aquatic environment. As EDCs, they can cause serious impacts on aquatic organisms as well as human health [77, 78]. Phthalates can combine with hormone receptors in the body to cause nervous system disorders, endocrine disorders, and immune decline, resulting in reproductive and developmental damage [79]. In a study by [80], they assessed the occurrence and evaluation of ecological risks of phthalates in the Persian Gulf and reported algae, crustaceans, and fish were at high risk of the chemical. In a separate experiment, [81] assessed the fate, toxicity, and ecological risk of phthalates using aquatic and terrestrial microcosms. They reported that both water and sediment media are at risk of phthalates and derived guidelines for their protection. This outcome agrees with [82] who reported impacts on toxicity and bioaccumulation of phthalates in organisms. Environmental levels of DBP have been assessed around the world. In many cases, levels found were lower than the highest levels assessed in this study [83, 84, 85], but in some cases, environmental DBP levels exceeded the levels assessed in this research [86, 87, 88]. These references report on DBP levels in freshwater in many countries, including India, Spain, France, Korea, China, Canada, the Netherlands, the United Kingdom, and South Africa. Of these, high levels of DBP were most commonly reported in South Africa [86, 88, 89]. Higher levels of DBP were widely associated with discharge from wastewater treatment works [88, 90]. Although the dose–response curve fitting procedures' results were not highly significant, the apparent response, particularly of lipid peroxidase, to DBP is of concern given the high levels reported from South African rivers.

The inhibition of AChE and the increase in LPx observed in C. nilotica and T. sparrmanii upon exposure to CAS, DBP, and BPA reveal critical mechanisms of toxicity. AChE inhibition indicates neurophysiological disruption, while elevated LPx reflects oxidative damage to cellular membranes. In South African freshwater ecosystems, where aquatic organisms are already subject to multiple anthropogenic pressures, such impairments could diminish species survival, disrupt ecological interactions, and compromise ecosystem functioning. These findings highlight the potential for plasticizers to exacerbate existing environmental stressors, threatening the biodiversity and sustainability of vital South African freshwater systems.

5. Conclusion

The test plasticizers (BPA, CAS, and DBP) assessed in this study were all generally toxic to the test organisms ( T. sparrmanii and C. nilotica ) after short‐term and long‐term exposures. Where mortality and biochemical effects occurred, DBP seemed to be the most harmful, followed by BPA and then CAS, which did not cause any harm to the tested organisms in most cases. It appears that the potential toxicity of CAS to freshwater organisms is limited by its solubility, but more of a threat is posed by BPA and DBP, which can induce short‐term mortality at levels where they are soluble in water. However, toxicity is likely to be limited by the rapid degradation of these plasticizers under natural conditions. More importantly, this study evaluated the ecotoxicology of BPA, CAS, and DBP in South African freshwater ecosystems for the first time by assessing organismal and biochemical effects on T. sparrmanii and C. nilotica as test organisms. The use of AChE and lipid peroxidase activities as endpoint biomarkers in this experiment provides valuable insights into neurotoxic effects and oxidative stress induced by plasticizers. These biomarkers are essential for evaluating the health status of T. sparrmanii , C. nilotica , and, by implication, other freshwater organisms in South Africa. As the concentrations used in the study were partly based on those that have been reported previously in aquatic ecosystems, this study is critical to understanding the broader ecological implications of plasticizer exposure. The outcome has provided important information for the protection of South African ecosystems against these toxicants.

Author Contributions

P.K.M. was involved with conceptualization, design, experimentation, data analyses, original draft preparation, review, and editing. N.M. supported with conducting the experiments and sample analyses. M.F.A.A. reviewed and edited the manuscripts. N.J.G. and O.N.O. were involved in data analyses and review. All authors have read and agreed to the published version of the manuscript.

Ethics Statement

The Rhodes University Animal Research Ethics Committee (RU‐AREC), which approves all animal experiments, approved the study (approval number 2021‐0684‐6375); the experiments were performed following institutional ethical guidelines.

Consent

The authors have nothing to report.

Conflicts of Interest

The authors declare no conflicts of interest.

Mensah P. K., Griffin N. J., Mgaba N., Akwetey M. F. A., and Odume O. N., “Toxicity of Commonly Used Plasticizers to the Freshwater Organisms Tilapia sparrmanii (Fish) and Caridina nilotica (Shrimp): Lethal and Sublethal Effects,” Environmental Toxicology 41, no. 1 (2026): 81–94, 10.1002/tox.24563.

Data Availability Statement

The data that support the findings of this study are available on request from the corresponding author. The data are not publicly available due to privacy or ethical restrictions.