Effects of Bisphenol Analogs on Marine Species: History
View Latest Version
Please note this is an old version of this entry, which may differ significantly from the current revision.
Contributor:
Valerio Matozzo
,
JACOPO FABRELLO

Bisphenol A analogs are currently used in manufacturing and as plasticizers as a substitute for bisphenol A. This replacement is taking place because bisphenol A is recognized as an endocrine disruptor chemical (EDC) that can also cause oxidative stress and genotoxic effects in aquatic species. Bisphenol A analogs have a similar chemical structure to  bisphenol A (BPA), raising doubts about their use as safer substitutes.

  • Bisphenol A analogs
  • marine invertebrates

1. Effects of Bisphenol A

The effects of bisphenols have been evaluated on some marine species, such as microalgae, mollusks, rotifers, sea urchins, polychaetae, crustaceans, fishes, and mammals. However, bisphenol A (BPA) is still the most studied bisphenol, while the effects of the other compounds are substantially unknown. BPA is considered as an endocrine disruptor chemical (EDC), being able to mimic the natural estrogens causing impairment in hormonal sensitivity and responsiveness. It is well known that BPA can also affect the immune system, antioxidant enzymes, neuroendocrine system, and embryo development in humans [1]. Similarly, several studies demonstrated that BPA could cause detrimental effects on marine species. BPA can alter the reproductive system of marine species, as demonstrated in Mytilus edulis specimens exposed for 3 weeks to 50 μg/L of bisphenol A [2]. The authors also observed a slight increase in phospho-proteins in the mantle gonadal tissue of females, and after histological investigations, they observed atretic oocytes in half of the BPA-exposed mantles, while on the other half there were post-spawning stage gonads [2]. Effects on the reproductive system were also reported in the mud crab Macrophthalmus japonicus where a significant upregulation of the vitellogenin (VTG) gene was observed in the ovaries after 96 h of exposure to the tested concentration (1, 10, and 30 µg/L) and a significant upregulation of VTG gene was also observed in hepatopancreas at 30 µg/L. Moreover, a significantly higher VTG gene expression was observed after 7 days of exposure at 1 µg/L in hepatopancreas and under all the tested concentrations in ovaries [3]. The same authors investigated the effect of BPA on the molting process. After one day of exposure, there was an ecdysone receptor gene up-regulation (EcR) in the hepatopancreas at 1, 10, and 30 µg/L, while in gills only 1 and 10 µg/L caused a significant up-regulation. However, the EcR gene expression in hepatopancreas was downregulated at 1 and 30 µg/L after 4 days. Interestingly, after one week of exposure, an opposite trend between the two tissues was observed, with the 10 µg/L treatment that caused a significant decrease in expression in gills and a significant increase in hepatopancreas [4].

In another crustacean species, namely, the whiteleg shrimp Litopenaeus vannamei, exposure to 2 µg/L of BPA induced a significantly smaller gonad-somatic index with a consequent delay in the gonad development stage with respect to the controls. Moreover, exposed shrimps had a lower oxygen consumption rate, an increased ammonia extraction rate, and a downregulation of metabolism-related gene expression. In addition, the authors observed an upregulation of gonadal development-related hormones and the expression of gene-encoding regulatory hormones [5]. It was also highlighted that BPA can cause embryotoxic effects [6][7][8]. For instance, in the mussel Mytilus galloprovincialis, BPA interfered with shell formation at different larval stages with spatial alteration of the expression of genes involved in shell formation and in serotoninergic system development [9]. In the same species, Balbi et al. [10] observed several gene expression alterations in embryos hatched from eggs previously exposed to 10 μg/L of BPA. In a similar study, in which fertilized eggs of Haliotis diversicolor supertexta were exposed to four BPA concentrations (0.05, 0.2, 2, and 10 μg/mL), it was demonstrated that BPA can affect embryonic development. Moreover, the authors concluded that BPA could markedly reduce embryo hatchability, increase developmental malformation, and suppress the metamorphosis behavior of larvae [11]. Larvae malformations were also recorded in the two ascidian species Ciona robusta and Ciona intestinalis after exposure to concentrations higher than 10 μM [8]. Furthermore, most of the embryos of the sea urchin Hemicentrotus pulcherrimus exposed to 10 μM of BPA for 24 or 48 h after fertilization showed a suppressed development by the hatching stage [6]. In addition, juveniles of H. pulcherrimus exposed for 80 days to 0.5 µM of BPA showed a reduction in the relative test diameter [6]. However, the authors reported that BPA effects on early development were less remarkable than that of ethynyl estradiol [6]. An analog study reported that BPA can affect the development of embryos of the sea urchin Paracentrotus lividus. Indeed, animals exposed to 300 µg/L of BPA showed spermiotoxic and embryotoxic effects and skeleton malformation was observed in plutei [7]. Moreover, larvae skeletal malformations were also observed in the embryos of sea urchin Arbacia lixula after BPA exposure [12]. Lastly, fertilized eggs of the sea urchin Lytechinus pictus exposed to BPA showed failed cytokinesis leading to multipolar spindles in a dose-dependent manner [13].

BPA can also affect the immune system in bivalves. Indeed, BPA injected in M. galloprovincialis (25 nM nominal concentration in the hemolymph) caused a significant lysosomal membrane destabilization in hemocytes at all the post-injection times (6, 12, and 24 h). Moreover, BPA changed the phosphorylation state of mitogen-activated protein kinases (MAPKs) and signal transducers and activators of transcription (STAT), indicating that BPA can affect kinase-mediated cell signaling in mussel hemocytes in vivo [14]. Furthermore, in the bivalve Tegillarca granosa total hemocyte count (THC) was reduced after 2 weeks of exposure to 10 ng/L and 100 ng/L of BPA and a decrease in red granulocyte percentage and an increase in both basophil granulocyte and hyalinocyte was reported [15]. In addition, the phagocytic activities of hemocytes were significantly reduced and the content of γ-aminobutyric acid, dopamine, and acetylcholine in hemolymph was increased [15]. On the contrary, the expression of four immune-related genes and genes encoding modulatory enzymes and receptors for neurotransmitters was significantly suppressed [15]. An impairment of the neuro system was also observed in the claw muscles of the artic spider crab Hyas araneus in which there was a significant reduction in acetylcholinesterase activity (AChE) after 3 weeks of exposure to 50 µg/L of BPA [16]. In another study, the crab Charybdis japonica was exposed for 1, 3, 6, 9, and 15 days to 0.125, 0.25, 0.5, and 1 mg/L, respectively. The authors reported a reduction in THC values in crabs exposed to 1 mg/L for 1, 3, and 6 days; 6 days of exposure to 0.5 mg/L of BPA also caused a THC reduction [17]. They also observed that superoxide dismutase (SOD), catalase (CAT), glutathione peroxidase (GPx), lysozyme (LSZ), and phenoloxidase (PO) activities reached the highest values during the first week and then decreased during the second week of exposure in both hemolymph and hepatopancreas, while malondialdehyde content (MDA) gradually increased over time following the exposure [17]. Histological analysis revealed that the rough endoplasmic reticulum of hepatopancreas cells appeared swollen and reduced in number [17]. Moreover, vacuoles, lysosomes, and myeloid bodies were observed in the hepatopancreas cells [17]. In addition, in the same crab species, there was a significantly increased expression of the heat shock protein gene HSP90 under all the tested concentrations (0.05, 0.5, 1 mg/L) after 12, 24, 48, and 96 h of exposure [18].

BPA can cause detrimental effects on the antioxidant system also in bivalves. Indeed, in specimens of M. galloprovincialis injected with 50 μL of BPA solutions (from a 10 mM stock solution in ethanol diluted in artificial seawater), containing respectively, 3, 15, and 60 ng BPA, corresponding to a nominal concentration of BPA 3, 15, and 60 ng/g dry weight or per mussel, BPA caused an increased gene expression of the estrogen receptor MeER2 and induced downregulation of antioxidant genes, catalase, and metallothioneins 24 h post injection [19]. In addition, BPA altered the activity of CAT, glutathione S-transferase (GST), glutathione reductase (GR), and the total glutathione amount [19]. Moreover, exposure for 7 days of Perna viridis to 98, 996, and 10,111 ng/L of BPA had immunomodulatory, genotoxic, and endocrine-disruptive effects [20].
The effects of BPA were also evaluated in polychaetae. In Ophryotrocha diadema BPA caused a significant reduction in the number of laid eggs only after five weeks of exposure at the highest concentration tested (1.4611 mg/L) [21]. Furthermore, in the polychaete Perinereis aibuhitensis exposed for 4, 7, and 14 days to 10, 50, and 100 µg/L, respectively, BPA caused the increase in G protein alpha subunit gene expression in both the body wall and in the head, on which there was the higher induced expression [22]. Lastly, recent studies have shown that some degradation products and metabolites of BPA have much higher estrogenicity or toxicity than BPA. For instance, 4-methyl-2,4-bis(4-hydroxyphenyl)pent-1-ene (MBP), a metabolite of bisphenol A, has shown an estrogenic activity approximately 1000-fold higher than BPA [23][24].

2. Effects of Bisphenol A Analogs

The effects of BPA analogs are poorly studied in marine species. However, in a recent study, bisphenol F (BPF), bisphenol S (BPS), and BPA were tested in the marine rotifer Brachionus koreanus for 24 h [25]. The authors reported that both BPA and BPF caused a reduction in cumulative offspring. In addition, 10 mg/L of BPA, 10 mg/L of BPF, and 15 mg/L of BPF caused a significant reduction in life span. Moreover, the three bisphenols increased the reactive oxygen species level (ROS), with BPS and BPF increasing the ROS and GST levels at almost all the tested concentrations (1, 5, and 10 mg/L) [25]. BPF as well as BPA significantly altered the expression level of cytochrome P450 (CYP) and GST genes [25].
Furthermore, in the brackish water flea Diaphanosoma celebensis, the gene expression of seven ecdysteroid pathway-related genes (cyp314a1, EcRA, EcRB, USP, nvd, HR3, and E75) were altered after exposure for 48 h at high concentrations of BPA (0.12, 0.6, and 3.0 mg/L), BPS (0.92, 4.6, and 23.0 mg/L), and BPF (0.6, 1.0, and 5.0 mg/L), suggesting an ecdysteroid signaling pathway disruption and an alteration of the endocrine system. However, the expression patterns of BPS and BPF were different from those of BPA [26][27]. In the same species exposed to the same concentrations, BPA, BPS, and BPF differently modulated the gene expression of the estrogen-related receptors, vitellogenin and vitellogenin receptors, indicating that these compounds can also affect the normal reproduction-related pathway [28]. Moreover, an in silico study on the Pacific oyster Magallana gigas revealed that BPF has a high affinity for the estrogen receptor (ER) and that BPA has a higher binding energy for ER than the estrogen hormone itself [29]. In a recent in vitro study, TBBPA, BPA-E (Bisphenol A BIS (2,3-dihydroxy propyl) ether), and bisphenol AF (BPAF) tested at 50 µM caused the reduction in residual carboxylesterase activity in the digestive gland extracts of both Octopus vulgaris and Sepia officinalis [30]. Similarly, the three BPA analogs caused an inhibition effect also in the hemolymph of O. vulgaris [30]. Recently, the effects of several bisphenol A analogs on β-esterases were also tested in the sea turtle Caretta caretta. The results showed that only TBBPA (at 50 μM) significantly inhibit the plasmatic carboxyl esterase activity, however, TBBPA did not inhibit acetylcholinesterase [31]. Furthermore, in the marine amphipod Gammarus aequicauda exposed to 0.25, 0.5, and 1 mg/L of BPA, BPF, or BPS for 24 h, there was a general increase in DNA damage in both hemocytes and spermatozoa. In detail, BPF caused a significant increase in DNA damage at all the tested concentrations in hemocytes and in spermatozoa at 1 mg/L, while the BPS exposure increased the mean DNA damage level with respect to the controls in both somatic and germ cells, but not significantly. The authors concluded that both BPF and BPS caused lower DNA damage than BPA [32].
In a recent study, juveniles of the brown trout Salmo trutta were exposed for 2 or 8 weeks to 2 or 20 mg/kg fish. After 2 weeks, the level of the thyroid hormone triiodothyronine (T3) in plasma was elevated after Bisphenol S exposure at the high concentration and paralleled by an increase in micronucleated cells. BPS did not cause statistical differences in the hematocrit, hemoglobin levels, or glucose levels in comparison with the control. However, there was a significantly higher hematocrit level in fish exposed to the high dose of BPS compared to fish exposed to the low dose of BPS after 2 weeks of exposure. On the contrary, the vitellogenin levels in blood were not altered by BPS, and only the higher dose of BPA increased the level. After 2 weeks, T3 levels were significantly higher in fish treated with the high dose of BPS compared with controls, but thyroxin T4 levels were not altered. At the same time as the exposure, the high dose of BPS increased the micronuclei percentage in the erythrocytes, while after 8 weeks, the same treatment decreased the percentage of binucleated erythrocytes [33]. Furthermore, the hepatocytes of the rainbow trout Oncorhyncus mykiss treated for 24 h with BPS (0, 15.63, 31.25, 62.50, 125, 250, and 500 µM) showed that cytotoxicity increased in a concentration-dependent manner. Moreover, all the tested BPS concentrations caused a reduction in SOD activity, while CAT and GPX activity was generally increased at higher concentrations. GST activity was significantly increased at a concentration of 31.25 µM or higher, while GST Theta 1-1 activity and the reduced glutathione content (GSH) were decreased at these concentrations. Moreover, the oxidative damage measured as malondialdehyde content increased at 125, 250, and 500 µM of BPS [34]. In an analog study, the hepatocytes of rainbow trout were exposed to BPF using the same experimental design. As in the case of BPS, BPF increased dose-dependently its cytotoxic effects. The malondialdehyde content was increased at BPF concentrations between 15.63 and 250 µM, whereas it remained unchanged at 500 µM. Interestingly, SOD and CAT activities were increased and decreased, respectively, in all treatment concentrations. Moreover, the GSH level increased with concentrations of BPF between 15.63 and 250 µM but decreased significantly at 500 µM. In addition, GPX and GST were significantly increased at a BPF concentration from 31.25 µM, and at 125 µM, respectively. The authors concluded that the toxic mechanism of BPF was mainly based on cell membrane damage and oxidative stress, influencing the antioxidant defenses [35]. The effect of BPS was also tested on the juveniles of olive flounders (Paralichthys olivaceus) that were injected with a concentration of 50 mg/kg. Treatment caused a transcriptome alteration in the liver. In particular, BPS significantly increased the transcription of egg process and vitellogenesis-related genes, including zona pellucida sperm-binding proteins, and estrogen receptors, with increases in plasma 17β-estradiol (E2) and VTG concentrations. In addition, there was an increased gene expression of genes involved in antioxidant defense systems, while genes involved in innate immunity were decreased [36]. Moreover, there was an increased activity of both CAT and GST in the liver [36].
Recently, specimens of medaka (Oryzias melastigma) were exposed for 70 days to 200 µg/L of BPA, BPF, BPAF, or their mixture [37]. After 70 days of exposure to BPAF, males showed a higher body weight and body length. On the other side, the condition factor of males was significantly reduced by the mixture and increased in females owing to BPF exposure [37]. In addition, the BPAF-exposed fishes had a survival rate significantly lower than controls. Moreover, the histological analysis indicated that bisphenol exposure led to vacuolization and local lesions in the liver, especially in the BPAF group. These results are like those observed by Peng et al. [17] in crabs exposed to BPA. In addition, medaka fishes showed an altered antioxidant enzyme activity, with a reduction in both SOD and CAT activity in the liver of males exposed to BPAF or MIX. In females, both BPF and BPAF caused a reduction in the total swimming distance, while in males this reduction was observed only in the mixed treatment. All the bisphenols caused an up-regulation of genes involved in lipid synthesis and metabolism in the liver of female fish and, interestingly, all genes involved in eukaryotic ribosome biogenesis were downregulated in females exposed to BPAF [37]. In males, the differentially expressed genes were mainly involved in steroid biosynthesis, arachidonic acid metabolism, nicotinate, and nicotinamide metabolism [37]. Moreover, BPAF appeared to cause an estrogenic effect with an increased coriogenins and vitellogenins gene expression in the liver of male fishes [37].

As a concluding remark, it can be highlighted that BPA analogs can act as EDCs and can also affect the immune system, antioxidant system, genetics, and behavior in aquatic species. BPA is going to be replaced with structurally similar compounds that are speculatively considered safer compounds. In the context of higher BPA analog usage and release into aquatic environments, recent studies have highlighted that the marine concentrations can already pose environmental risks to non-target species. However, the effects of BPA analogs and their mixtures on marine organisms are still mainly unknown and need to be deeply investigated.

This entry is adapted from the peer-reviewed paper 10.3390/jmse10091271

References

  1. Ma, Y.; Liu, H.; Wu, J.; Yuan, L.; Wang, Y.; Du, X.; Wang, R.; Marwa, P.W.; Petlulu, P.; Chen, X.; et al. The adverse health effects of bisphenol A and related toxicity mechanisms. Environ. Res. 2019, 176, 108575.
  2. Aarab, N.; Lemaire-Gony, S.; Unruh, E.; Hansen, P.D.; Larsen, B.K.; Andersen, O.K.; Narbonne, J.F. Preliminary study of responses in mussel (Mytilus edilus) exposed to bisphenol A, diallyl phthalate and tetrabromodiphenyl ether. Aquat. Toxicol. 2006, 78, S86–S92.
  3. Park, K.; Jo, H.; Kim, D.K.; Kwak, I.S. Environmental pollutants impair transcriptional regulation of the vitellogenin gene in the burrowing mud crab (Macrophthalmus japonicus). Appl. Sci. 2019, 9, 1401.
  4. Park, K.; Kim, W.S.; Kwak, I.S. Characterization and transcriptional response of ecdysone receptor gene in the mud crab Macrophthalmus japonicus: Effects of osmotic stress and endocrine disrupting chemicals. Ocean. Sci. J. 2019, 54, 611–620.
  5. Han, Y.; Shi, W.; Tang, Y.; Zhou, W.; Sun, H.; Zhang, J.; Yan, M.; Hu, L.; Liu, G. Microplastics and bisphenol A hamper gonadal development of whiteleg shrimp (Litopenaeus vannamei) by interfering with metabolism and disrupting hormone regulation. Sci. Total Environ. 2022, 810, 152354.
  6. Kiyomoto, M.; Kikuchi, A.; Unuma, T.; Yokota, Y. Effects of ethynylestradiol and bisphenol A on the development of sea urchin embryos and juveniles. Mar. Biol. 2006, 149, 57–63.
  7. Özlem, Ç.A.; Hatice, P. Effects of bisphenol A on the embryonic development of sea urchin (Paracentrotus lividus). Environ. Toxicol. 2008, 23, 387–392.
  8. Mercurio, S.; Messinetti, S.; Barzaghi, B.; Pennati, R. Comparing the sensitivity of two cogeneric ascidian species to two plastic additives: Bisphenol A and the flame retardant tris (chloro-propyl) phosphate. Eur. Zool. J. 2022, 89, 437–445.
  9. Miglioli, A.; Balbi, T.; Besnardeau, L.; Dumollard, R.; Canesi, L. 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.
  10. Balbi, T.; Franzellitti, S.; Fabbri, R.; Montagna, M.; Fabbri, E.; Canesi, L. Impact of bisphenol A (BPA) on early embryo development in the marine mussel Mytilus galloprovincialis: Effects on gene transcription. Environ. Pollut. 2016, 218, 996–1004.
  11. Zhou, J.; Zhu, X.S.; Cai, Z.H. The impacts of bisphenol A (BPA) on abalone (Haliotis diversicolor supertexta) embryonic development. Chemosphere 2011, 82, 443–450.
  12. Arslan, O.C.; Parlak, H. Effects of bisphenol-A on the embryological development of the sea urchin Arbacia Lixula (Linnaeus, 1758). Fresenius Environ. Bull. 2008, 17, 127.
  13. George, O.; Bryant, B.K.; Chinnasamy, R.; Corona, C.; Arterburn, J.B.; Shuster, C.B. Bisphenol A directly targets tubulin to disrupt spindle organization in embryonic and somatic cells. ACS Chem. Biol. 2008, 3, 167–179.
  14. Canesi, L.; Betti, M.; Lorusso, L.C.; Ciacci, C.; Gallo, G. ‘In vivo’ effects of Bisphenol A in Mytilus hemocytes: Modulation of kinase-mediated signalling pathways. Aquat. Toxicol. 2005, 71, 73–84.
  15. Tang, Y.; Zhou, W.; Sun, S.; Du, X.; Han, Y.; Shi, W.; Liu, G. Immunotoxicity and neurotoxicity of bisphenol A and microplastics alone or in combination to a bivalve species, Tegillarca granosa. Environ. Pollut. 2020, 265, 115115.
  16. Minier, C.; Forget-Leray, J.; Bjørnstad, A.; Camus, L. Multixenobiotic resistance, acetyl-choline esterase activity and total oxyradical scavenging capacity of the Arctic spider crab, Hyas araneus, following exposure to bisphenol A, tetra bromo diphenyl ether and diallyl phthalate. Mar. Pollut. Bull. 2008, 56, 1410–1415.
  17. Peng, Y.Q.; Wang, M.J.; Chen, H.G.; Chen, J.H.; Gao, H.; Huang, H.H. Immunological responses in haemolymph and histologic changes in the hepatopancreas of Charybdis japonica (A. Milne-Edwards, 1861) (Decapoda: Brachyura: Portunidae) exposed to bisphenol A. J. Crustac. Biol. 2018, 38, 489–496.
  18. Park, K.; Kwak, I.S. Characterize and gene expression of heat shock protein 90 in marine crab Charybdis japonica following bisphenol A and 4-nonylphenol exposures. Environ. Health Toxicol. 2014, 29, e2014002.
  19. Canesi, L.; Borghi, C.; Ciacci, C.; Fabbri, R.; Vergani, L.; Gallo, G. Bisphenol-A alters gene expression and functional parameters in molluscan hepatopancreas. Mol. Cell. Endocrinol. 2007, 276, 36–44.
  20. Juhel, G.; Bayen, S.; Goh, C.; Lee, W.K.; Kelly, B.C. Use of a suite of biomarkers to assess the effects of carbamazepine, bisphenol A, atrazine, and their mixtures on green mussels, Perna viridis. Environ. Toxicol. Chem. 2017, 36, 429–441.
  21. Ruberto, S.; Buono, D.; Santovito, A. Polychaetes as bioindicators of environmental Pollution: Impact of bisphenol A on the reproduction rate of Ophryotrocha diadema (Åkesson, 1976) (eunicida: Dorvilleidae). Zool. Ecol. 2021, 31, 61–65.
  22. Zhao, H.; Zhou, Y.; Li, Y.; Li, S.; Yang, D. Molecular cloning and expression of the gene for G protein alpha subunit induced by bisphenol A in marine polychaete Perinereis aibuhitensis. Environ. Toxicol. Pharmacol. 2014, 37, 521–528.
  23. Okuda, K.; Takiguchi, M.; Yoshihara, S. In vivo estrogenic potential of 4- methyl-2,4-bis(4-hydroxyphenyl)pent-1-ene, an active metabolite of bisphenol A, in uterus of ovariectomized rat. Toxicol. Lett. 2010, 197, 7–11.
  24. Yoshihara, S.; Mizutare, T.; Makishima, M.; Suzuki, N.; Fujimoto, N.; Igarashi, K.; Ohta, S. Potent estrogenic metabolites of bisphenol A and bisphenol B formed by rat liver S9 fraction: Their structures and estrogenic potency. Toxicol. Sci. 2004, 78, 50–59.
  25. Park, J.C.; Lee, M.C.; Yoon, D.S.; Han, J.; Kim, M.; Hwang, U.K.; Jung, J.H.; Lee, J.S. Effects of bisphenol A and its analogs bisphenol F and S on life parameters, antioxidant system, and response of defensome in the marine rotifer Brachionus koreanus. Aquat. Toxicol. 2018, 199, 21–29.
  26. In, S.; Yoon, H.W.; Yoo, J.W.; Cho, H.; Kim, R.O.; Lee, Y.M. Acute toxicity of bisphenol A and its structural analogues and transcriptional modulation of the ecdysone-mediated pathway in the brackish water flea Diaphanosoma celebensis. Ecotoxicol. Environ. Saf. 2019, 179, 310–317.
  27. In, S.; Cho, H.; Lee, Y.M. Identification of ecdysteroid pathway-related genes and their transcriptional modulation in the brackish water flea Diaphanosoma celebensis exposed to bisphenol analogs. Toxicol. Environ. Health. Sci. 2021, 13, 261–268.
  28. In, S.; Cho, H.; Lee, K.W.; Won, E.J.; Lee, Y.M. Cloning and molecular characterization of estrogen-related receptor (ERR) and vitellogenin genes in the brackish water flea Diaphanosoma celebensis exposed to bisphenol A and its structural analogues. Mar. Pollut. Bull. 2020, 154, 111063.
  29. Raj, A.; Nair, S.N.; Abdulvahab, R.; Ittoop, G. In-Silico Modelling of Interaction Between Environmental Xenoestrogens and Estrogen Receptor of Pacific Oyster (Magallana gigas ) Using AutoDock. Inform. Stud. 2022, 9.
  30. Omedes, S.; Andrade, M.; Escolar, O.; Villanueva, R.; Freitas, R.; Solé, M. B-Esterases characterisation in the digestive tract of the common octopus and the European cuttlefish and their in vitro responses to contaminants of environmental concern. Environ. Res. 2022, 210, 112961.
  31. Sole, M.; Bassols, A.; Labrada-Martagón, V. Plasmatic B-esterases as potential biomarkers of exposure to marine plastics in loggerhead turtles. Environ. Res. 2022, 213, 113639.
  32. Cosentino, S.; Aureli, F.; Iannilli, V. Bisphenols A and its analogues induce genotoxic damage in marine and freshwater amphipods. Environ. Adv. 2022, 7, 100183.
  33. Frenzilli, G.; Martorell-Ribera, J.; Bernardeschi, M.; Scarcelli, V.; Jönsson, E.; Diano, N.; Moggio, M.; Guidi, P.; Sturve, J.; Asker, N. Bisphenol A and bisphenol S induce endocrine and chromosomal alterations in brown trout. Front. Endocrinol. 2021, 12, 645519.
  34. Kaptaner, B.; Yılmaz, C.; Aykut, H.; Doğan, E.; Fidan, C.; Bostancı, M.; Yıldız, F. Bisphenol S leads to cytotoxicity-induced antioxidant responses and oxidative stress in isolated rainbow trout (Oncorhyncus mykiss) hepatocytes. Mol. Biol. Rep. 2021, 48, 7657–7666.
  35. Aykut, H.; Kaptaner, B. In vitro effects of bisphenol F on antioxidant system indicators in the isolated hepatocytes of rainbow trout (Oncorhyncus mykiss). Mol. Biol. Rep. 2021, 48, 2591–2599.
  36. Jung, J.H.; Moon, Y.S.; Kim, B.M.; Lee, Y.M.; Kim, M.; Rhee, J.S. Comparative analysis of distinctive transcriptome profiles with biochemical evidence in bisphenol S-and benzo pyrene-exposed liver tissues of the olive flounder Paralichthys olivaceus. PLoS ONE 2018, 13, e0196425.
  37. Li, X.; Liu, Y.; Chen, Y.; Song, X.; Chen, X.; Zhang, N.; Li, H.; Guo, Y.; Wang, Z.; Dong, Z. Long-Term exposure to bisphenol A and its analogues alters the behavior of marine medaka (Oryzias melastigma) and causes hepatic injury. Sci. Total Environ. 2022, 841, 156590.
More
© Text is available under the terms and conditions of the Creative Commons Attribution (CC BY) license; additional terms may apply. By using this site, you agree to the Terms and Conditions and Privacy Policy.
This entry is offline, you can click here to edit this entry!
ScholarVision Creations
Feedback