Abstract

Bisphenol E (BPE) has emerged as a potential environmental contaminant similar to its analogue Bisphenol A (BPA). This study aimed to evaluate the ecotoxicological effects of BPE on two marine model organisms, Phaeodactylum tricornutum and Mytilus galloprovincialis. The algal growth inhibition test revealed an EC₅₀ value of 1.94 mg L⁻¹ after 72 h, indicating significant growth inhibition at higher concentrations. In the acute toxicity test with M. galloprovincialis, the LC₅₀ value was 5.391 mg L⁻¹ after 96 h. Extended exposure (14 days) to sublethal concentrations resulted in decreased Condition Index (CI) and Gonadosomatic Index (GSI), with CI values dropping from 16% in the control to 5% at 32 mg L⁻¹, and GSI decreasing from 68% to 30%. In vitro fertilization assays demonstrated that high concentrations of BPE (1000 ng L⁻¹) significantly reduced fertilization success. These findings suggest that BPE poses substantial risks to marine ecosystems, affecting both primary producers and higher trophic-level species.

Keywords

Bisphenol E, Mytilus galloprovincialis, Phaeodactylum tricornutum, Ecotoxicology, Physiological effects, Algal growth inhibitions. 

Highlights

• It was determined that Bisphenol E inhibited the growth of Phaeodactylum tricornutum species by causing inhibition (EC₅₀: 1,94 mg L^-1).
• It was determined that bisphenol E had a physiological effect on the Mytilus galloprovincialis species (LC₅₀: 5.391 mg L ^-1), causing a decrease in CI% and GSI% due to an increase in concentration.
• Contrary to the first two concentrations (0.01-0.1 ng L^-1), it was determined that the fertilization success of mussels was negatively affected at increasing concentrations.
• Both model organisms were observed to be highly sensitive.

Introduction

Bisphenol A (BPA) is a diphenylmethane derivative containing two hydroxyphenyl groups. It is a monomer used in the production of polycarbonate plastics and epoxy resins and is also a raw material in the manufacturing of numerous products, including thermal papers, plastic containers, toys, medical equipment, dental filling materials, cigarette filters, and the internal coatings of metal food cans.^1-6 Due to its high production and usage, this chemical inevitably enters the environment and has been detected in various environmental compartments such as water, soil, sediment, indoor dust, and human tissues.^7,8 Numerous studies over the years have shown that BPA, besides being an endocrine disruptor, can cause adverse health effects such as cardiovascular diseases, obesity, diabetes, reproductive disorders, and breast cancer.^9-11 Following the discovery of its risks and adverse effects on both humans and other living organisms, the production and usage of Bisphenol A have been restricted. For instance, in 2010, its use in the manufacture of baby bottles was banned in Canada.¹² Following these various restrictions and bans, the production of alternative chemicals to replace BPA began. These new chemicals, known as BPA analogues, consist of two phenol groups connected by chemical structures similar to BPA. According to the ECHA¹³ report, there are 17 identified compounds with a Bisphenol structure. These analogues include Bisphenol AF (BPAF), Bisphenol AP (BPAP), Bisphenol B (BPB), Bisphenol BP (BPBP), Bisphenol C (BPC), Bisphenol F (BPF), Bisphenol E (BPE), Bisphenol G (BPG), Bisphenol M (BPM), Bisphenol P (BPP), Bisphenol S (BPS), Bisphenol Z (BPZ), Bisphenol A diglycidyl ether (BADGE), Bisphenol PH (BPPH), Bisphenol FL (BPFL), Bisphenol TMC (BPTMC).

Despite being known for their use in the production of various products, the overall production and usage quantities of all analogues are not yet fully known. Many analogues act as polymers during production, suggesting that their production quantities are likely higher than estimated or detectable values. Research has determined that the concentrations of these analogues in aquatic ecosystems vary between µg/L and ng/L.¹² The average concentration range for Bisphenol E has been reported as 40.6 ± 36.6 ng L^-1.¹⁴ The entry of all analogues, including Bisphenol E, into aquatic environments occurs through inefficient removal during the production, treatment, and processing of bisphenols in wastewater treatment facilities, leakage from regular storage sites, and household solid waste.^15-17 Research indicates that BPs, especially BPA, are released from plastics that come into contact with food and products intended for children, such as food storage containers, water bottles, baby bottles, cups, and toys. For example, Wang et al.¹⁸ investigated the release of BPs from polycarbonate and polyethylene terephthalate bottles into water. The average concentrations of certain bisphenols like BPA, BPE, and BPAF in bottled water were found to be 20.8, 1.8, and 2.2 ng L^-1, respectively. Literature reviews have shown that the focus of studies has been on analogues such as BPS, BPF, and BPAF. In the study by dos Santos et al., 2024,¹⁴ the acute risk was found to be high for BPA and BPAF (algae, crustaceans, and fish), BPF (fish and bacteria), and BPS (algae). For BPB (algae, crustaceans, and fish), BPF (algae and crustaceans), BPS (crustaceans and bacteria), and BPZ (fish), the risks were considered medium. Conversely, it was noted that there was no information provided for BPE due to a lack of data. The insufficient number of studies on BPE, the absence of environmental risk assessments, and the uncertainty of its potential effects led to the decision to investigate this analogue in the study.

Primary producers in aquatic ecosystems, such as phytoplanktonic organisms, consist of several taxonomic groups, including cyanobacteria, diatoms, and green algae. These photosynthetic microorganisms play a crucial role in mediating carbon, nutrient (N and P), and oxygen biogeochemical cycles in aquatic ecosystems. Besides their significant contribution as primary producers in the food web,¹⁹ they can also induce substantial adverse effects on other organisms at higher trophic levels. Therefore, they are considered as foundational organisms in ecotoxicological studies, utilized to determine the initial response to the effects of new chemicals. While numerous studies have investigated the effects of Bisphenol A on phytoplanktonic organisms, limited studies are available regarding the effects of known analogues.

Bivalve species, due to their broad geographical distribution, sessile lifestyle, and limited mobility, reflect the conditions of their habitats and have been termed 'sentinel species' since 1975.²⁰ The advantages of using bivalves as sentinel species include their presence in globally extensive and accessible populations, high sensitivity to specific water pollutants, the ability to biologically accumulate various xenobiotics, their significant role in the food chain, and their capability to assimilate hydrophilic/lipophilic chemicals.^21-23 The utilization of mussels as protectors against environmental pollution has extended to pharmaceuticals²², microplastics,²⁴ and other emerging pollutants such as endocrine-disrupting compounds.²⁵ Specifically, for determining the effects of endocrine-disrupting compounds and monitoring their environmental status, bivalves are considered model organisms as sentinel species.²⁶

Understanding the physiological effects of chemicals on Mytilus galloprovincialis species serves a paramount purpose in environmental research. It is crucial to discern how various chemical compounds impact the physiology of this species as it plays a fundamental role in marine ecosystems. Investigating these effects aids in assessing the potential risks posed by pollutants, encompassing their impacts on physiological parameters such as condition index, gonadosomatic index, and reproductive success. This research is vital for comprehending the broader implications of chemical contamination on marine biodiversity and ecosystem dynamics. Additionally, studying these effects provides valuable insights into the health and resilience of marine organisms, helping to gauge the overall ecological health of aquatic environments.

İn this context, the ecotoxicological effects of Bisphenol E, one of the analogues produced as an alternative to the endocrine-disrupting compound Bisphenol A, were assessed on two species, Phaeodactylum tricornutum and Mytilus galloprovincialis, which hold significant positions in the aquatic ecosystem across trophic levels.

Materials and Methods

Bisphenol E (BPE; C₁₄H₁₄O₂) purchased from Sigma-Aldrich 98% (Cas. No: 2081-08-5, Molecular weight: 214.26 g/mol) was dissolved in pure water. The stock solutions were stored in dark vials at 4°C. Intermediate stocks were prepared and diluted to the required concentrations for the experiments.

Test Species and Culture Condition 
Phaeodactylum tricornutum

Phaeodactylum tricornutum was selected as the marine species for the assays due to its widespread presence as a phytoplankton in marine environments. The algae stock provided for the propagation and environment of pure culture was cultivated in sterilized (autoclave, 121ºC 15 min) natural seawater and regularly enriched with F/2 medium.²⁷ Silica (SiO₂, 50μg/L), Nitrate (NO_3-, 6 μg/L) and Phosphate (PO₄, 6 μg/L) were also added to the medium to prevent the possibility of chelators or trace elements interacting with substances in the sediment.²⁸ The ambient temperature was kept constant at 20ºC and illumination was achieved with cold-white light (100 μmol photon/m²/s). P.tricornutum culture was kept under these laboratory conditions for 3 months before the study was started. 

Mytilus galloprovincialis
Mussels were collected from a clean area that is not exposed to any domestic or industrial wastewater, in the İzmir-Çeşme (Aegean coast of Turkey 38°27′ 00.01 N″–26°37′ 27.60 E″). Before the experiment, all mussels were taken from a stock tank (260×80×70 cm³). They were acclimatized (one week) in the laboratory in tanks of artificial seawater (After purchasing Natural Sourching sea salt cas number: 7647-14-5, This product does not meet the definition of a hazardous substance or preparation as defined by the European Union Council Directives 67/548/EEC, 1999/45/EC, 1272/2008/EC and subsequent Directives) in the laboratory environment) (34.5 ± 0.2 psu, 5.1 ± 0.1 mg L^-1 dissolved oxygen, and 8.1 ± 0.1 pH) areated continuously at 17.5±1°C. In the experiment, 50x30x30 cm³ and 20 L volume glass aquariums were used. Mussels were fed daily with the phytoplankton Chlorella sp.

Experimental Setup
Algal Growth Inhibition Test

The algal growth inhibition test was conducted following the OECD. Preliminary range-finding tests were conducted to identify a suitable concentration range that would elicit sublethal to moderate toxic effects in Phaeodactylum tricornutum. Based on these trials, the selected nominal concentrations for the definitive bioassays were 0.5, 0.8, 1.0, 1.5, and 2.0 mg L⁻¹ for BPE. No analytical verification of actual concentrations was performed; thus, all values reported in this study are nominal. The decision not to conduct chemical analysis was due to resource limitations, and this is acknowledged as a limitation of the study. However, the selected concentrations align with or fall within the upper range of those previously reported in environmental monitoring studies (e.g., Russo et al., 2021).¹⁴ During the bioassay stage, BPE was added to the medium containing 20 mL of nutrient solution (F/2) at different concentrations (0.5-0.8-1.0-1.5, and 2.0 mg L^-1 each). All test concentrations were performed in 6 replicates and cell counts were performed by visual counting under a microscope at 0, 24, 48 and 72 hours. The determined growth curves were compared with the control group and the inhibition percentage was calculated.

Acute Toxicity Test with Mussels
The acute toxicity test was performed according to the OECD guideline 203 and Parlak et al. 2011.²⁹ After the acclimatization period, the initial step was the conduction of an acute toxicity test. Artificial seawater was introduced into glass aquariums in the test setup, followed by the installation of aeration. Subsequently, 20 mussels were placed in each glass aquarium. Throughout the entire trial, the temperature was maintained at 16°C, while monitoring oxygen and pH levels to ensure optimal conditions. Similar to the algal test, the selected exposure concentrations (2-4-6-8-16-32 mg L^-1) for Mytilus galloprovincialis were determined through preliminary range-finding tests. These concentrations were chosen to capture a dose-response curve up to significant mortality levels. No chemical analysis was conducted to confirm the actual concentrations in the solutions; therefore, the data presented are based on nominal concentrations. The mussels were exposed to higher concentrations of BPE (2-4-8-16-32 mg L^-1 per aquarium) in 20 L glass aquariums containing 20 mussels each. The exposure time was 96 hours and mortality was recorded at 24-hour intervals. The criterion used to distinguish between dead and surviving mussels was the degree of shell opening; mussels with completely open shells were deemed dead due to the opposing force between their muscles and ligaments. 

Physiological Measurements with Mussels
Following the acute toxicity test, surviving mussels (n=8) were exposed to BPE (2-4-8-16-32 mg L^-1) for 14 days to assess physiological effects. At the end of the exposure period, mussels were dissected to measure physiological parameters such as Condition Index (CI) and Gonadosomatic Index (GSI). The CI was calculated as the ratio of dry flesh weight to shell weight, while the GSI was determined as the ratio of gonad weight to total soft tissue weight. The impact of Bisphenol E on the Condition Index (CI) was determined following the criteria by Matozzo et al.³⁰ Soft tissues were collected by separating the shells from four individuals. Both shells and soft tissues were dried in an oven at 60°C for 48 hours. The Gonadosomatic Index (GSI) was determined according to Peters and Granek.³¹ Dissection was performed on four individuals, and gonads were separated from all tissues and dried in an oven at 60°C for 48 hours. 

Fertilization Success with Mussels
To evaluate the effect of BPE on reproductive success, in vitro fertilization assays were conducted. The mussels brought into the laboratory environment were individually placed in glass jars. The water temperature was naturally raised to induce the release of sperm and eggs, ranging from 16°C to 22°C. Subsequently, the release of eggs and sperm was observed. After identifying male and female individuals distinguished by a change in water color, eggs and sperm were respectively filtered through 100 µm and 55 µm filters. Following the selection of three healthy male and female individuals based on criteria set by His et al.³² test concentrations (0.01-0.1-1-100-1000 ng L^-1) were chosen according to the calculated LC50 values. Stock solutions prepared for each chemical were diluted and used accordingly. In a 6-well plate, 9 ml of artificial seawater and increasing chemical concentrations were added, followed by an egg:sperm ratio of 1:10. After 30 minutes, fertilization success was fixed with 4% formalin for microscopic observation. Fertilized eggs were distinguished by the formation of the fertilization membrane.

Statistical Analysis

Growth inhibition rates were calculated according to OECD (201)³³ guidelines using the following formula:
μ₀–ⱼ = (ln xⱼ − ln x₀)/(tⱼ − t₀)

Percentage inhibition (Ir %) for each test concentration was then determined as:
Ir % = [(μc − μr)/μc] × 100

IC₅₀ values were derived using the “area under the curve” method, and comparisons with the control were performed by Dunnett’s test (p < 0.05). LC₅₀ values for mussel assays were calculated via EPA probit analysis software (v1.5). CI and GSI were computed using the standard formulas, and fertilization ratios were evaluated by one-way ANOVA followed by Dunnett’s test (p < 0.05). All analyses were carried out with Statistica 12.0.

Results

Algal Growth Inhibition Test
The potential effects of Bisphenol E (BPE) on the growth rate of Phaeodactylum tricornutum were investigated using an algal growth inhibition test. The test was conducted according to OECD guidelines, and preliminary tests were carried out to determine the appropriate concentration range. The algal growth inhibition test results indicated that the control group exhibited a 1.5-fold increase in phytoplankton cell count over 72 hours, indicating normal growth and no restrictive effects. The test concentrations (0.5-0.8-1.0-1.5-2.0 mg L^-1) were compared with the control group.

The data showed that at the lowest concentration (0.5 mg L^-1), there was no significant inhibition of algal growth. However, at a concentration of 0.8 mg L^-1, a slight inhibition was observed. At 1.0 mg L^-1, the inhibition effect decreased, but as the concentration increased to 1.5 mg L^-1 and 2.0 mg L^-1, significant inhibition effects were observed, culminating in a 50% inhibition at the highest concentration (2.0 mg L^-1). The EC₅₀ value for BPE was calculated to be 1.94 mg L^-1 after 72 hours (Figure 1).

 

Acute Toxicity Test with Mussels
The acute toxicity of BPE on Mytilus galloprovincialis was assessed following OECD guidelines. Mussels were exposed to increasing concentrations of BPE (2-4-8-16-32 mg L^-1) for 96 hours. Mortality rates were recorded at 24-hour intervals, and the LC₅₀ value was determined to be 5.391 mg L^-1.

Physiological Measurements with Mussels
Surviving mussels from the acute toxicity test were exposed to BPE concentrations (2-4-8-16-32 mg L^-1) for an extended period of 14 days to evaluate the effects on physiological parameters. The Condition Index (CI) and Gonadosomatic Index (GSI) were measured. The results showed a decrease in CI values compared to the control group. The CI decreased from 8% at 2 mg L^-1 to 5% at 32 mg L^-1, while the control group maintained a CI of 16%. Despite this decreasing trend, no significant differences were found among the various concentrations (Figure 2). Similarly, the GSI decreased from 68% at 2 mg L^-1 to 30% at 32 mg L^-1, with significant differences observed among the concentrations (Figure 3).

 

 

Fertilization Success with Mussels
The effect of BPE on the fertilization success of Mytilus galloprovincialis was evaluated by conducting in vitro fertilization assays. Mussels were exposed to BPE concentrations (0.01-0.1-1-100-1000 ng L^-1). No effect was observed on fertilization success at the lowest concentrations (0.01 and 0.1 ng L^-1). However, at higher concentrations, BPE negatively impacted fertilization success, with the most significant reduction observed at 1000 ng L^-1. The detailed percentages and corresponding DE values are presented in Table 1 and Figure 4.

 

 

 

Discussion

The study comprehensively examined the ecotoxicological responses of P. tricornutum and M. galloprovincialis to Bisphenol E exposure. Compared with previously reported BPA toxicity (EC₅₀ = 0.6 mg L⁻¹ in P. tricornutum),³⁴ BPE exhibited a moderate yet measurable inhibitory effect (EC₅₀ = 1.94 mg L⁻¹). The observed response aligns with the compound’s lipophilic character (high log Kₒw), suggesting cell membrane interactions as a primary mechanism of toxicity.³⁵

In mussels, sublethal exposure caused notable reductions in physiological indices (CI and GSI), consistent with disrupted energy allocation between somatic growth and reproduction. Extended exposure appeared to trigger energy redirection toward defense and detoxification pathways.^36-39

Fertilization assays further demonstrated BPE’s sublethal reproductive effects, showing no change at 0.01–0.1 ng L⁻¹ but significant reduction at 100–1000 ng L⁻¹. These findings are comparable to the reproductive inhibition previously observed with BPA, suggesting that BPE may elicit similar endocrine-disrupting mechanisms.^40-44

Risk Quotient (RQ) analysis indicated high ecological risk (RQ > 1) for both test organisms, underscoring the need for monitoring BPE in aquatic environments.^45-48

Environmental Risk Assessment
Based on EC50 and LC50 data together with the maximum environmental concentrations (MECs) reported in surface waters, we calculated Risk Quotient (RQ) values to evaluate the potential ecological risks of BPE. The RQ was determined using the standard formula (RQ = MEC / PNEC) and classified according to Czarny-Krzymińska et al. (2023), where an RQ < 0.01 indicates an insignificant risk, 0.01–0.1 low, 0.1–1 medium, and >1 high risk. The results revealed that BPE poses a high risk for both test organisms: M. galloprovincialis (RQ = 1.99) and P. tricornutum (RQ = 4.58). These findings highlight the urgent need for continuous environmental monitoring and stricter regulatory measures for BPE, as well as further studies to better understand its long-term effects on aquatic organisms and ecosystem health.

Conclusion

This study provides important insights into the ecotoxicological effects of Bisphenol E (BPE) on marine organisms. The findings demonstrate that BPE causes significant sublethal effects, including growth inhibition in P. tricornutum and physiological and reproductive impairments in M. galloprovincialis. The high RQ values calculated for both species further confirm the considerable ecological risks posed by BPE in aquatic environments. These results emphasize the urgent need for stricter monitoring and regulation of BPE and its analogues, along with more comprehensive studies focusing on their long-term impacts on marine ecosystems. Overall, this research contributes to the growing evidence that emerging BPA analogues can pose serious threats to aquatic biodiversity and ecosystem balance.

Acknowledgement

It is part of the doctoral thesis numbered 843656, registered with the Turkish Council of Higher Education. The present study was supported Ege University Faculty of Fisheries, Hydrobiology Department (Project No: 22892). 

References

1. Adhikary A. Bisphenol A: Prenatal Exposure and Its Effect on Obesity and Male Reproductive System. In Proceedings of the Zoological Society. December 2021; New Delhi: Springer India.
2. Pelch K, Wignall JA, Goldstone AE, et al. A scoping review of the health and toxicological activity of bisphenol A (BPA) structural analogues and functional alternatives. Toxicology. 2019;424:152235.
3. Tarafdar A, Sirohi R, Balakumaran PA, et al The hazardous threat of Bisphenol A: Toxicity, detection and remediation. Journal of Hazardous Materials. 2022;423:127097.
4. Vaccher V, Lopez ME, Castaño A, et al. European interlaboratory comparison investigations (ICI) and external quality assurance schemes (EQUAS) for the analysis of bisphenol A, S and F in human urine: Results from the HBM4EU project. Environmental Research. 2022;210:112933.
5. Zhou R, Xia M, Zhang L, et al. Individual and combined effects of BPA, BPS and BPAF on the cardiomyocyte differentiation of embryonic stem cells. Ecotoxicol Environ Saf. 2021;220:112366.
6. Xing J, Zhang S, Zhang M, et al. A critical review of presence, removal and potential impacts of endocrine disruptors bisphenol A. Comp Biochem Physiol C Toxicol Pharmacol. 2022;254:109275.
7. Kang JH, Kondo F, Katayama Y. Human exposure to bisphenol A. Toxicology. 2006;226(2-3):79-89.
8. Roark AM. Endocrine Disruptors and Marine Systems. In: Goldstein MI, Della Sala DA, Eds., Encyclopedia of the World’s Biomes, Elsevier, Oxford, 2020;pp.188-194.
9. Giulivo M, Lopez de Alda M, Capri E, et al. Human exposure to endocrine disrupting compounds: Their role in reproductive systems, metabolic syndrome and breast cancer. A review. Environ Res. 2016;151:251-264. 
10. Rochester JR. Bisphenol A and human health: a review of the literature. Reprod Toxicol. 2013;42:132-155.
11. Ziv-Gal A, Flaws JA. Evidence for bisphenol A-induced female infertility: a review (2007-2016). Fertil Steril. 2016;106(4):827-856. 
12. Czarny-Krzymińska K, Krawczyk B, Szczukocki D. Bisphenol A and its substitutes in the aquatic environment: Occurrence and toxicity assessment. Chemosphere. 2023;315:137763.
13. ECHA. Assessment of Regulatory Needs. 
14. dos Santos CR, Arcanjo GS, Araújo AAD, et al. Occurrence, environmental risks, and removal of bisphenol A and its analogues by membrane bioreactors. Chemical Engineering Journal. 2024;494:153278.
15. Godiya C, Park B. Removal of bisphenol A from wastewater by physical, chemical and biological remediation techniques. A review. Environ Chem Lett. 2022;20:1801-1837s.
16. Huang Z, Zhao JL, Zhang CY, et al. Profile and removal of bisphenol analogues in hospital wastewater, landfill leachate, and municipal wastewater in South China. Sci Total Environ. 2021;790:148269. 
17. Liu J, Zhang L, Lu G, et al. Occurrence, toxicity and ecological risk of Bisphenol A analogues in aquatic environment - A review. Ecotoxicol Environ Saf. 2021;208:111481. 
18. Wang H, Liu ZH, Tang Z, et al. Bisphenol analogues in Chinese bottled water: Quantification and potential risk analysis. Sci Total Environ. 2020;713:136583.
19. Paerl H, Justic D. Primary producers: phytoplankton ecology and trophic dynamics in coastal waters. In: Wolanski E, McLusky DS. Eds., Treatise on Estuarine and Coastal Science, 2011;6:23-42s.
20. Farrington JW, Tripp BW, Tanabe S, et al. Edward D. Goldberg's proposal of “the mussel watch”: reflections after 40 years. Marine Pollution Bulletin. 2016;110(1):501-510s.
21. Chiesa LM, Nobile M, Malandra R, et al. Occurrence of antibiotics in mussels and clams from various FAO areas. Food Chemistry. 2018;240:16-23s.
22. Świacka K, Maculewicz J, Smolarz K, et al. Mytilidae as model organisms in the marine ecotoxicology of pharmaceuticals - A review. Environ Pollut. 2019;254(Pt B):113082. 
23. Staniszewska M, Graca B, Sokołowski A, et al. Factors determining accumulation of bisphenol A and alkylphenols at a low trophic level as exemplified by mussels Mytilus trossulus. Environ Pollut. 2017;220(Pt B):1147-1159.
24. Li J, Lusher AL, Rotchell JM, et al. Using mussel as a global bioindicator of coastal microplastic pollution. Environ Pollut. 2019;244:522-533.
25. Chiu JM, Po BH, Degger N, et al. Contamination and risk implications of endocrine disrupting chemicals along the coastline of China: a systematic study using mussels and semipermeable membrane devices. Science of the total environment. 2018;624:1298-1307s.
26. Omar TFT, Aris AZ, Yusoff FM, et al. Occurrence and level of emerging organic contaminant in fish and mollusk from Klang River estuary, Malaysia and assessment on human health risk. Environ Pollut. 2019;248:763-773. 
27. Guıllard RR, Ryther JH. Studies of marine planktonic diatoms. I. Cyclotella nana Hustedt, and Detonula confervacea (cleve) Gran. Can J Microbiol. 1962;8:229-239.
28. Moreno-Garrido I, Hampel M, Lubián LM, et al. Marine Benthic Microalgae Cylindrotheca Closterium (Ehremberg) Lewin and Reimann (Bacillariophyceae) as a Tool for Measuring Toxicity of Linear Alkylbenzene Sulfonate in Sediments. Bulletin of environmental contamination and toxicology. 2003;70:242-247.
29. OECD. Test No. 203: Fish, Acute Toxicity Test, OECD Guidelines for the Testing of Chemicals, Section 2, OECD Publishing, Paris; 2019.
30. Matozzo V, Formenti A, Donadello G, et al. A multi-biomarker approach to assess effects of Triclosan in the clam Ruditapes philippinarum. Mar Environ Res. 2012;74:40-46. 
31. Peters JR, Granek EF. Long-term exposure to fluoxetine reduces growth and reproductive potential in the dominant rocky intertidal mussel, Mytilus californianus. Science of the Total Environment. 2016;545:621-628s.
32. His E, Beiras R, Seaman MNL. The assessment of marine pollution-bioassays with bivalve embryos and larvae. In Advances in marine biology. Academic press. 1999;37:1-178.
33. OECD (Organization for Economic Cooperation and Development), Guideline for Testing of Chemicals no. 201. Alga Growth Inhibition Test, Paris; 1984.
34. Seoane M, Cid Á, Esperanza M. Toxicity of bisphenol A on marine microalgae: Single- and multispecies bioassays based on equivalent initial cell biovolume. Sci Total Environ. 2021;767:144363.
35. Li R, Liu Y, Chen G, et al. Physiological responses of the alga Cyclotella caspia to bisphenol A exposure. Bot Mar. 2008;51(5):360–369s.
36. Czarny K, Krawczyk B, Szczukocki D. Toxic effects of bisphenol A and its analogues on cyanobacteria Anabaena variabilis and Microcystis aeruginosa. Chemosphere. 2021;263:128299.
37. Ding T, Li W, Yang M, et al. Toxicity and biotransformation of bisphenol S in freshwater green alga Chlorella vulgaris. Sci Total Environ. 2020;747:141144.
38. Ji MK, Kabra A, Choi J, et al. Biodegradation of bisphenol A by the freshwater microalgae Chlamydomonas mexicana and Chlorella vulgaris. Ecol Eng. 2014;73:260–269s.
39. Li J, Li W, Huang X, et al. Comparative study on the toxicity and removal of bisphenol S in two typical freshwater algae. Environ Sci Pollut Res Int. 2021;28(27):36861-36869.
40. Naveira C, Rodrigues N, Santos FS, et al. Acute toxicity of Bisphenol A (BPA) to tropical marine and estuarine species from different trophic groups. Environ Pollut. 2021;268(Pt B):115911.
41. Guo L, Li Z, Gao P, et al. Ecological risk assessment of bisphenol A in surface waters of China based on both traditional and reproductive endpoints. Chemosphere. 2015;139:133-137.
42. Chen M, Ike M, Fujita M. Acute toxicity, mutagenicity, and estrogenicity of bisphenol-A and other bisphenols. Eniron Toxicol. 2001;17(1):80-86.
43. Andral B, Stanisiere JY, Sauzade D, et al. Monitoring chemical contamination levels in the Mediterranean based on the use of mussel caging. Marine Pollution Bulletin. 2004;49(9-10):704-712s.
44. Oliveira P, Almeida Â, Calisto V, et al. Physiological and biochemical alterations induced in the mussel Mytilus galloprovincialis after short and long-term exposure to carbamazepine. Water Res. 2017;117:102-114. 
45. Miglioli A, Balbi T, Besnardeau L, et al. Bisphenol A interferes with first shell formation and development of the serotoninergic system in early larval stages of Mytilus galloprovincialis. Sci Total Environ. 2021;758:144003.
46. Ginebreda A, Muñoz I, de Alda ML, et al. Environmental risk assessment of pharmaceuticals in rivers: relationships between hazard indexes and aquatic macroinvertebrate diversity indexes in the Llobregat River (NE Spain). Environment international. 2010;36(2):153-162.
47. Han Y, Fei Y, Wang M, et al. Study on the Joint Toxicity of BPZ, BPS, BPC and BPF to Zebrafish. Molecules. 2021;26(14):4180.
48. Jin X, Zha J, Xu Y, et al. Derivation of predicted no effect concentrations (PNEC) for 2,4,6-trichlorophenol based on Chinese resident species. Chemosphere. 2012;86(1):17-23.