were among the main findings. Effect concentrations were in the low µg/L range (high variation in sensitivity between species, studies and laboratories) and embryos being more sensitive than adults, but the precise mechanism is not clarified yet. The effects could be highly dependent on life stages, exposure window, and specific concentrations at specific time points which is already well-documented for the endocrine disruptive effects of tributyltin (TBT) on marine mollusks. *Marisa cornuarietis* and *Potamopyrgus antipodarum*, two prosobranch snails, were the most extensively studied. Since 2017, only a few *in vivo* single exposure BPA studies in mollusks have been published. In the gastropod *Pomacea lineata*, Andrade *et al.* (2017) (de Andrade *et al.*, 2017) found decreased neonate heart rate, adult behavioral changes and 96-h LC₅₀ values of 3.14 mg/L (neonate) and 11.1 mg/L (adults) which is comparable with earlier gastropod studies (Mihaich *et al.*, 2009). In bivalves, BPA exposure caused impaired antimicrobial ability (Tang *et al.*, 2022) and female-biased sex ratio (Lusher *et al.*, 2017).

The literature concerning toxic or endocrine disruptive effects of BPA alternatives in mollusk *in vivo* studies is limited to a few reports on the effects of TBBPA. The effects of TBBPA were investigated and compared between several vertebrate and invertebrate species including two bivalve mollusks: the Eastern oyster (*Crassostrea virginica*) and the blue mussel (*Mytilus edulis*). Mollusca was the most sensitive aquatic phylum and shell growth was the most sensitive endpoint in both short-term (96-h, *C. virginica*) and long-term (70-d, *M. edulis*) exposures with NOEC and LOEC values in *M. edulis* of 0.017 and 0.032 mg/L, respectively (reviewed by Pittinger & Pecquet, 2018 (Pittinger and Pecquet, 2018).

The fact that BPA and BPA alternatives have been detected in mollusks in the environment (Baralla et al., 2021) shows the importance of including mollusks in further studies of BPA and its alternatives. Especially early life stages of mollusks seem to be sensitive towards BPA exposure, but the effects of BPA alternatives, except TBBPA, are not investigated in mollusks. *L. stagnalis* (adults and embryos) as a model species appears to be appropriate due to the promising toxicity testing methods currently under development. In *L. stagnalis* (adults and embryos) neither BPA nor its alternatives have been tested and an evaluation of the impact of BPA and its alternatives in *L. stagnalis* (adults and embryos) should be initiated by studies of BPA itself and followed by single and combined exposures with BPA alternative.

### 4.5.2.3. Ascidian tunicates.
The vase tunicate (*Ciona intestinalis*), which is an invasive ascidian tunicate, has been important model organisms in developmental biology and genomic evolution research for more than a century, due to their phylogenetic position as the earliest chordates, and the potential for using ascidian tunicates as ED bioindicators has gained increasing attention lately (Gomes et al., 2019; Capela et al., 2020; Morthorst et al., 2023), and the topic was recently discussed (Beyer et al., 191 (2023)). The increasing interest on *Ciona* (and related ascidians) is also due to their wide distribution and abundant presence in shallow coastal seas, uncomplicated maintenance in laboratory cultures, rapid embryo development and short life cycle, well characterized developmental ontology, tadpole-like swimming larvae, the transparent body of the sessile adults, and their fully sequenced genome. With tunicates being the closest living invertebrate relatives to all vertebrates, the effects caused by EDCs on these test models could hypothetically have broad relevance, maybe in some cases for both invertebrate and chordate taxa. However, the degree by which Ciona/ascidians are usable as bioindicators in ED assessments relies on the amount of knowledge that is available on their endocrinology, their sensitivity and responsiveness to known EDCs, as well as the degree to which their putative ED responses are comparable to other taxonomic groups. These are topics for which there are still many knowledge gaps. But notwithstanding these challenges, using ascidians as test species for assessing ED properties of BPA and its substitutes could be an option worth considering, and this constitutes the scope of this chapter.

About 40 types of hormones/neuropeptides have been identified and

characterized in ascidians by purification, cDNA cloning, and peptidomic approaches, and the main ascidian hormones/peptides that have been found can be classified into three categories: (1) vertebrate homologs, such as gonadotropin-releasing hormones, tachykinins, galanin-like peptides, calcitonin and insulin/relaxin paralogs; (2) hormones/peptides that belong to conserved families but with distinct sequences or activities compared to vertebrate homologs, such as GnRH-X and vasopressin; and (3) hormones/peptides that are ascidian-specific, such as Ci-YFV/Ls (novel neuropeptides specific to *Ciona* with unknown biological functions but which share a Tyr-Phe-Val/Leu sequence at the C-terminus) and Ci-LFs (eight novel neuropeptides possibly specific to *Ciona* that share a Lys-Phe sequence at the C-terminus), see multiple references in Beyer *et al.* 2023 (*Beyer et al.*, 191 (2023)). However, the function of many of these hormones/peptides remain uncharacterized. Nuclear Receptors (NRs) that are positively identified in ascidians could represent important focus points for studies into possible ED effect phenomena and ED related MOAs in ascidian taxa. Studies of the sequenced genome of *Ciona* have identified at least 17 genes which encode for NR transcription factors (Yagi *et al.*, 2003). NR genes documented to be present in ascidians include, among others, the thyroid receptor (TR), the peroxisome proliferator-activated receptor (PPAR), the retinoic acid/retinoid X receptors (RAR/RXR), the vitamin D/pregnane X receptors (VDR/PXR), and the estrogen-related receptor (ERR), whereas NR related genes that are present in humans, but thought NOT to be present in ascidians, include most importantly the steroid hormones and their receptors, i.e., the estrogen receptor (ER), androgen receptor (AR), mineralocorticoid receptor (MR), glucocorticoid receptor (GR), and progesterone receptor (PR), for references to original studies see Beyer *et al.* 2023 (*Beyer et al.*, 191 (2023)).

BPA is possibly the ED substance (beside tributyltin, TBT) which so far have been most studied in ascidians, although the total number of studies is still quite limited, i.e., (Messinetti *et al.*, 2019; Messinetti *et al.*, 2018; Cangialosi *et al.*, 2013; Matsushima *et al.*, 2013; Mansueto *et al.*, 2011; Mercurio *et al.*, 2022; Richter and Fidler, 2015; Gomes *et al.*, 2019; Ferrari *et al.*, 2022; Gazo *et al.*, 2021).

To highlight some of these studies, Messinetti *et al.* (2018; 2019) (Messinetti *et al.*, 2019; Messinetti *et al.*, 2018) investigated dose-responsive effects of BPA on ontogenetic development in larvae of two ascidian species, *Ciona* and the related ascidian *Phallusia mammillata*, identifying specific disorders (in neural differentiation and swimming tail and pigmented organs morphology) as markers of ED effects, and suggested also which NR system (Estrogen-related receptor, ERR) and which target cells (GABAergic and dopaminergic neurons) that were affected by the BPA exposure. Involvement of upregulation of ERR activity was suggested as MOA as co-exposure of ascidians with 4-hydroxytyramoxifen, a compound that binds and deactivates ERRs, resulted in normal pigmented-organ phenotypes. Gomes *et al.* (2019) (Gomes *et al.*, 2019) addressed the role of ERR in neurodevelopmental toxicity of bisphenol A to embryos of the ascidian *Phallusia mammillata*. Cangialosi *et al.*, (2013) (Cangialosi *et al.*, 2013) assessed the effect of BPA and atrazine in *Ciona* embryos and found concentration-dependent effects on embryo development (development arrest, fused or absent blastomeres and lack of follicular and test cells). Matsushima *et al.* (2013) (Matsushima *et al.*, 2013), investigated how low (1 microM) BPA exposure of early (neurula stage) embryos adversely affected the swimming behavior of the subsequent larva. The study suggested a mechanism that involved BPA binding to and disrupting the regulatory action of ERR receptors, although the molecular mechanism was not clarified. Mansueto *et al.* (2011) (Mansueto *et al.*, 2011) used post-larvae juvenile *Ciona* as test models for detecting impacts of BPA (and TBT) on organ morphology developments and found dose-responsive effects on the morphology of tunic, gonad cells, nervous system, digestive system, as well as and inhibition effect of the rhythmic body contractions. Mercurio *et al.* (2022) (Mercurio *et al.*, 2022) studied teratogenic effects of BPA and the flame retardant tris(chloro-propyl) phosphate on embryo development in *Ciona*. Lastly, Richter and Fidler (2015) (Richter and
were among the main findings. Effect concentrations were in the low µg/L range (high variation in sensitivity between species, studies and laboratories) and embryos being more sensitive than adults, but the precise mechanism is not clarified yet. The effects could be highly dependent on life stages, exposure window, and specific concentrations at specific time points which is already well-documented for the endocrine disruptive effects of tributyltin (TBT) on marine mollusks. *Marisa cornuarietis* and *Potamopyrgus antipodarum*, two prosobranch snails, were the most extensively studied. Since 2017, only a few *in vivo* single exposure BPA studies in mollusks have been published. In the gastropod *Pomacea lineata*, Andrade et al. (2017) (de Andrade et al., 2017) found decreased neonate heart rate, adult behavioral changes and 96-h LC₅₀ values of 3.14 mg/L (neonate) and 11.1 mg/L (adults) which is comparable with earlier gastropod studies (Mihaich et al., 2009). In bivalves, BPA exposure caused impaired antimicrobial ability (Tang et al., 2022) and female-biased sex ratio (Lusher et al., 2017).

The literature concerning toxic or endocrine disruptive effects of BPA alternatives in mollusk *in vivo* studies is limited to a few reports on the effects of TBBPA. The effects of TBBPA were investigated and compared between several vertebrate and invertebrate species including two bivalve mollusks: the Eastern oyster (*Crassostrea virginica*) and the blue mussel (*Mytilus edulis*). Mollusca was the most sensitive aquatic phylum and shell growth was the most sensitive endpoint in both short-term (96-h, *C. virginica*) and long-term (70-d, *M. edulis*) exposures with NOEC and LOEC values in *M. edulis* of 0.017 and 0.032 mg/L, respectively (reviewed by Pittinger & Pecquet, 2018 (Pittinger and Pecquet, 2018).

The fact that BPA and BPA alternatives have been detected in mollusks in the environment (Baralla et al., 2021) shows the importance of including mollusks in further studies of BPA and its alternatives. Especially early life stages of mollusks seem to be sensitive towards BPA exposure, but the effects of BPA alternatives, except TBBPA, are not investigated in mollusks. *L. stagnalis* (adults and embryos) as a model species appears to be appropriate due to the promising toxicity testing methods currently under development. In *L. stagnalis* (adults and embryos) neither BPA nor its alternatives have been tested and an evaluation of the impact of BPA and its alternatives in *L. stagnalis* (adults and embryos) should be initiated by studies of BPA itself and followed by single and combined exposures with BPA alternative.

### 4.5.2.3. Ascidian tunicates.
The vase tunicate (*Ciona intestinalis*), which is an invasive ascidian tunicate, has been important model organisms in developmental biology and genomic evolution research for more than a century, due to their phylogenetic position as the earliest chordates, and the potential for using ascidian tunicates as ED bioindicators has gained increasing attention lately (Gomes et al., 2019; Capela et al., 2020; Morthorst et al., 2023), and the topic was recently discussed (Beyer et al., 191 (2023)). The increasing interest on *Ciona* (and related ascidians) is also due to their wide distribution and abundant presence in shallow coastal seas, uncomplicated maintenance in laboratory cultures, rapid embryo development and short life cycle, well characterized developmental ontology, tadpole-like swimming larvae, the transparent body of the sessile adults, and their fully sequenced genome. With tunicates being the closest living invertebrate relatives to all vertebrates, the effects caused by EDCs on these test models could hypothetically have broad relevance, maybe in some cases for both invertebrate and chordate taxa. However, the degree by which *Ciona*/ascidians are usable as bioindicators in ED assessments relies on the amount of knowledge that is available on their endocrinology, their sensitivity and responsiveness to known EDCs, as well as the degree to which their putative ED responses are comparable to other taxonomic groups. These are topics for which there are still many knowledge gaps. But notwithstanding these challenges, using ascidians as test species for assessing ED properties of BPA and its substitutes could be an option worth considering, and this constitutes the scope of this chapter.

About 40 types of hormones/neuropeptides have been identified and

characterized in ascidians by purification, cDNA cloning, and peptidomic approaches, and the main ascidian hormones/peptides that have been found can be classified into three categories: (1) vertebrate homologs, such as gonadotropin-releasing hormones, tachykinins, galanin-like peptides, calcitonin and insulin/relaxin paralogs; (2) hormones/peptides that belong to conserved families but with distinct sequences or activities compared to vertebrate homologs, such as GnRH-X and vasopressin; and (3) hormones/peptides that are ascidian-specific, such as Ci-YFV/Ls (novel neuropeptides specific to *Ciona* with unknown biological functions but which share a Tyr-Phe-Val/Leu sequence at the C-terminus) and Ci-LFs (eight novel neuropeptides possibly specific to *Ciona* that share a Lys-Phe sequence at the C-terminus), see multiple references in Beyer et al. 2023 (Beyer et al., 191 (2023)). However, the function of many of these hormones/peptides remain uncharacterized. Nuclear Receptors (NRs) that are positively identified in ascidians could represent important focus points for studies into possible ED effect phenomena and ED related MOAs in ascidian taxa. Studies of the sequenced genome of *Ciona* have identified at least 17 genes which encode for NR transcription factors (Yagi et al., 2003). NR genes documented to be present in ascidians include, among others, the thyroid receptor (TR), the peroxisome proliferator-activated receptor (PPAR), the retinoic acid/retinoid X receptors (RAR/RXR), the vitamin D/pregnane X receptors (VDR/PXR), and the estrogen-related receptor (ERR), whereas NR related genes that are present in humans, but thought NOT to be present in ascidians, include most importantly the steroid hormones and their receptors, i.e., the estrogen receptor (ER), androgen receptor (AR), mineralocorticoid receptor (MR), glucocorticoid receptor (GR), and progesterone receptor (PR), for references to original studies see Beyer et al. 2023 (Beyer et al., 191 (2023)).

BPA is possibly the ED substance (beside tributyltin, TBT) which so far have been most studied in ascidians, although the total number of studies is still quite limited, i.e., (Messinetti et al., 2019; Messinetti et al., 2018; Cangialosi et al., 2013; Matsushima et al., 2013; Mansueto et al., 2011; Mercurio et al., 2022; Richter and Fidler, 2015; Gomes et al., 2019; Ferrari et al., 2022; Gazo et al., 2021).

To highlight some of these studies, Messinetti et al. (2018; 2019) (Messinetti et al., 2019; Messinetti et al., 2018) investigated dose-responsive effects of BPA on ontogenetic development in larvae of two ascidian species, *Ciona* and the related ascidian *Phallusia mammillata*, identifying specific disorders (in neural differentiation and swimming tail and pigmented organs morphology) as markers of ED effects, and suggested also which NR system (Estrogen-related receptor, ERR) and which target cells (GABAergic and dopaminergic neurons) that were affected by the BPA exposure. Involvement of upregulation of ERR activity was suggested as MOA as co-exposure of ascidians with 4-hydroxytyramoxifen, a compound that binds and deactivates ERRs, resulted in normal pigmented-organ phenotypes. Gomes et al. (2019) (Gomes et al., 2019) addressed the role of ERR in neurodevelopmental toxicity of bisphenol A to embryos of the ascidian *Phallusia mammillata*. Cangialosi et al., (2013) (Cangialosi et al., 2013) assessed the effect of BPA and atrazine in *Ciona* embryos and found concentration-dependent effects on embryo development (development arrest, fused or absent blastomeres and lack of follicular and test cells). Matsushima et al. (2013) (Matsushima et al., 2013), investigated how low (1 microM) BPA exposure of early (neurula stage) embryos adversely affected the swimming behavior of the subsequent larva. The study suggested a mechanism that involved BPA binding to and disrupting the regulatory action of ERR receptors, although the molecular mechanism was not clarified. Mansueto et al. (2011) (Mansueto et al., 2011) used post-larvae juvenile *Ciona* as test models for detecting impacts of BPA (and TBT) on organ morphology developments and found dose-responsive effects on the morphology of tunic, gonad cells, nervous system, digestive system, as well as and inhibition effect of the rhythmic body contractions. Mercurio et al. (2022) (Mercurio et al., 2022) studied teratogenic effects of BPA and the flame retardant tris(chloro-propyl) phosphate on embryo development in *Ciona*. Lastly, Richter and Fidler (2015) (Richter and
Fidler, 2015) explored the potential of reporter gene bioassays de-
velopments based on ascidian NR genes (ascidian VDR/PXRα LBDs as
fusion proteins combined with GAL4-DBD, a generic transcription acti-
vation domain (VP16-AD) and the lacZ reporter gene) as high-
throughput test tools for emerging substances with ED mechanisms
similar to BPA.

*Ciona intestinalis* is considered a promising model test species for
developing new approach methodologies (NAMs) for ED related testing
and screening, both regarding BPA substitutes and other chemicals of
emerging concern. As tunicates are the most primitive chordates alive,
such NAMs will meet the requirements on chordate test animal ethics.
However, considerably more research/knowledge is required, both on
the endocrine system in *Ciona/ascidians* and how it will be interfered
with by different EDC classes, before the usefulness of *Ciona* as an ED
model test species can be assessed. Although many vertebrate hormone
receptor paralogs have been identified in tunicates, their ligand-binding
status is still largely unknown. Hence, it will still be premature to
conclude that *Ciona* can serve as an alternative test model for screening
BPA substitutes and other, possibly ED acting, emerging contaminants.

4.5.2.4. *Crustacea.* *D. magna* is a small freshwater crustacean that plays a crucial role in the food chain and in the freshwater ecosystem, as it consumes mainly green algae and simultaneously is the main food source for fish. Therefore, its reproduction or behavioral disorders may influence the balance of the aquatic ecosystem. Considering its common distribution, short life cycle, large reproduction capacity as well as stable genotype and sensitivity to different chemicals *D. magna* is used as a model organism in aquatic toxicity tests of different environmental pollutants (Chen et al., 2021). The influence of specific compounds on *D. magna* can be assessed based on the different behavioral and physiological responses, such as *e. g.* immobilization, mortality, moulting, feeding, swimming and reproduction (Liu et al., 2019). In general, the reproduction in culture is leading to the production of females by diploid parthenogenesis. Males are only produced in response to environmental stress (e.g. high population densities). The body of a *Daphnia* is not clearly segmented, with the thoracic and abdominal regions covered in a secreted shell or carapace (Brennan et al., 2006).

Brennan et al. (2006) (Brennan et al., 2006) highlighted that in
crustaceans many physiological processes are regulated by neurohor-
mones, which includes lipid metabolism, digestion, ionic balance,
moulting, growth, regeneration, reproductive physiology, gonadal
development as well as cardiac activity (DeFur, 1999). Moreover, most
of the invertebrate hormones have been identified as peptide neuro-
hormones, however there has not been enough evidence on the inter-
ference of EDCs on peptide-base signaling systems. However, it has been
also highlighted that key insect and crustacean hormones are non-
peptide based, such as juvenile hormones or ecdysteroids, and pro-
cesses in which these hormones are involved may be more susceptible to
receptor related interference. For example, it has been hypothesized that
some chemicals which disrupt endocrine processes in vertebrates can
also negatively affect the hormonally related moulting process in ar-
thropods by acting as antagonists of endogenous ecdysteroids resulting
in the blocking (Zou and Fingerman, 1997) or activation of the ecdys-
teroid receptor (Song et al., 2017). The potential of chemically induced
molting interference has been the subject of several putative (concep-
tual) AOPs involving chitin synthesis and degradation (AOP #358, #359
and #361) and as well as one OECD endorsed AOP (AOP #360)
collectively organized under the OECD project 1.94 (https://www.aop
wiki.org). Other AOPs of relevance include ecdysone receptor-
mediated interference with molting (AOP #4) and interference with
male sexual development (AOP #201). Increasing attention is given
these toxicity mechanisms (Song et al., 2017; Schmid et al., 2023;
Schmid et al., 2021; Song et al., 2017; Toyota et al., 2022) in crustaceans
such as *D. magna*, although the number of chemicals tested remains low.
Nevertheless, the assessment of interference within the endocrine

system of invertebrates is difficult to predict. Despite these mentioned
difficulties with interpretation of the *Daphnia* tests' results for risk
assessment strategy of EDCs (Brennan et al., 2006); *D. magna* seems to be
a valuable model for evaluating endocrine disruptors' effects.

*D. magna* was also used for the acute and chronic toxicity assessment of BPA and its alternatives Supp.info 5 (Table 2). The available data for acute toxicity (including mainly standard endpoints such as immobilization or mortality) is also presented in the Supp.info 5 (Table 1). It might be concluded that most of the available data refers to BPA and the information about the toxicity of BPA alternatives is very inconsistent and limited. Even less data is available for the assessment of long-term effects of BPA alternatives towards *D. magna* (Supp.info 5 – Tab 1).

With respect to the prolonged exposure of *D. manga* to BPA and its alternatives, many different endpoints were investigated. This specifically includes not only reproductive indicators (such as: time to first brood, number of first brood per female, brood times, total number of offspring, number of offspring per brood per female and intrinsic rate of natural increase) as well as growth and development parameters (including swimming behavior, body length, heart rate and thoracic limb activity), but also biochemical indicators (acetyl choline esterase (AChE), superoxide dismutase (SOD), catalase (CAT), and glutathione S-transferase (GST) activities as well as malonic dialdehyde (MDA) content) (Liu et al., 2020).

Based on the available results it was confirmed that relative activity
of these enzymes in *D. magna* decreased after exposure to BPA, BPF, BPS
and their mixtures, implying that they may cause the disorder of
digestive, nervous and antioxidative system of *D. magna* (Liu et al.,
2019). Similar observation were also made in the case of BPF, which
induced the disturbance of neurological function and oxidative stress
situation in *D. magna*, leading to behavioral and reproductive effects,
such as growth inhibition, delay in sexual maturation, decreased
spawning, and eventually reduced population fertility (Liu et al., 2020).
Moreover, the observed correlations between biochemical, behavioral,
and reproductive endpoints indicated that the potential of biomarker
techniques as early predictors of bisphenol contaminants-induced al-
terations is also at a higher sensitivity level.

The available data presented in this chapter highlight the fact that
more comprehensive studies on the adverse effects of BPA and its al-
teratives on aquatic organisms are required at different levels of the
ecological hierarchical scale (Liu et al., 2020). Several marine crusta-
ceans have been proposed as potential candidates to investigate the
biological effects of contaminants on primary consumers. Brine shrimp
*Artemia* spp. (Crustacea, Branchiopoda, Anostraca), commonly known
as sea monkey, is a cosmopolitan organism inhabiting saline environ-
ments, such as inland saltwater lakes and coastal lagoons, feeding pri-
marily on phytoplankton and being an important primary consumer.
The main advantages of brine shrimp which make it a good test or-
ganism is the short life cycle thus a short test duration (i.e. 28–72 h from
hatching to the first endpoint) and the easy culturing, maintenance and
manipulation under laboratory conditions. Moreover, the use of nauplii
hatched from commercially available durable cysts (eggs) ensures
population homogeneity (organisms of similar age, genotype and
physiological condition) and availability all year-round without the
necessity of culturing (Nunes et al., 2006).

Artemia's sensitivity, like in most aquatic organisms, depends on its
developmental stage. According to literature, instar II-III (larval stage)
have been shown to be the most sensitive stage (Sorgeloos et al., 1978;
Vanhaecke and Persoone, 1984). One of the most important variables in
the performance of any biological experiment is the nutrition of the test
animals (Bengtson et al., 1984). In this context, one of the important
advantages of Artemia is that nauplii (stage II-III) do not require feeding
for up to 48 h. Therefore, in assessing the ecotoxicity of pollutants, the
acute toxicity test with mortality/immobilization (after 5–30 s obser-
vation) as the end point is most often used. Besides, sublethal indicators
of toxicity have already been applied for Artemia as biological endpoints,
such as biomarkers (acetylcholinesterase; heat shock proteins; lipid
Fidler, 2015) explored the potential of reporter gene bioassays de-
velopments based on ascidian NR genes (ascidian VDR/PXRα LBDs as
fusion proteins combined with GAL4-DBD, a generic transcription acti-
vation domain (VP16-AD) and the lacZ reporter gene) as high-
throughput test tools for emerging substances with ED mechanisms
similar to BPA.

Ciona intestinalis is considered a promising model test species for
developing new approach methodologies (NAMs) for ED related testing
and screening, both regarding BPA substitutes and other chemicals of
emerging concern. As tunicates are the most primitive chordates alive,
such NAMs will meet the requirements on chordate test animal ethics.
However, considerably more research/knowledge is required, both on
the endocrine system in Ciona/ascidians and how it will be interfered
with by different EDC classes, before the usefulness of Ciona as an ED
model test species can be assessed. Although many vertebrate hormone
receptor paralogs have been identified in tunicates, their ligand-binding
status is still largely unknown. Hence, it will still be premature to
conclude that Ciona can serve as an alternative test model for screening
BPA substitutes and other, possibly ED acting, emerging contaminants.

4.5.2.4. *Crustacea.* *D. magna* is a small freshwater crustacean that plays a crucial role in the food chain and in the freshwater ecosystem, as it consumes mainly green algae and simultaneously is the main food source for fish. Therefore, its reproduction or behavioral disorders may influence the balance of the aquatic ecosystem. Considering its common distribution, short life cycle, large reproduction capacity as well as stable genotype and sensitivity to different chemicals *D. magna* is used as a model organism in aquatic toxicity tests of different environmental pollutants (Chen et al., 2021). The influence of specific compounds on *D. magna* can be assessed based on the different behavioral and physiological responses, such as *e. g.* immobilization, mortality, moulting, feeding, swimming and reproduction (Liu et al., 2019). In general, the reproduction in culture is leading to the production of females by diploid parthenogenesis. Males are only produced in response to environmental stress (e.g. high population densities). The body of a *Daphnia* is not clearly segmented, with the thoracic and abdominal regions covered in a secreted shell or carapace (Brennan et al., 2006).

Brennan et al. (2006) (Brennan et al., 2006) highlighted that in
crustaceans many physiological processes are regulated by neurohor-
mones, which includes lipid metabolism, digestion, ionic balance,
moulting, growth, regeneration, reproductive physiology, gonadal
development as well as cardiac activity (DeFur, 1999). Moreover, most
of the invertebrate hormones have been identified as peptide neuro-
hormones, however there has not been enough evidence on the inter-
ference of EDCs on peptide-base signaling systems. However, it has been
also highlighted that key insect and crustacean hormones are non-
peptide based, such as juvenile hormones or ecdysteroids, and pro-
cesses in which these hormones are involved may be more susceptible to
receptor related interference. For example, it has been hypothesized that
some chemicals which disrupt endocrine processes in vertebrates can
also negatively affect the hormonally related moulting process in ar-
thropods by acting as antagonists of endogenous ecdysteroids resulting
in the blocking (Zou and Fingerman, 1997) or activation of the ecdys-
teroid receptor (Song et al., 2017). The potential of chemically induced
molting interference has been the subject of several putative (concep-
tual) AOPs involving chitin synthesis and degradation (AOP #358, #359
and #361) and as well as one OECD endorsed AOP (AOP #360)
collectively organized under the OECD project 1.94 (https://www.aop
wiki.org). Other AOPs of relevance include ecdysone receptor-
mediated interference with molting (AOP #4) and interference with
male sexual development (AOP #201). Increasing attention is given
these toxicity mechanisms (Song et al., 2017; Schmid et al., 2023;
Schmid et al., 2021; Song et al., 2017; Toyota et al., 2022) in crustaceans
such as *D. magna*, although the number of chemicals tested remains low.
Nevertheless, the assessment of interference within the endocrine

system of invertebrates is difficult to predict. Despite these mentioned
difficulties with interpretation of the *Daphnia* tests' results for risk
assessment strategy of EDCs (Brennan et al., 2006); *D. magna* seems to be
a valuable model for evaluating endocrine disruptors' effects.

*D. magna* was also used for the acute and chronic toxicity assessment of BPA and its alternatives Supp.info 5 (Table 2). The available data for acute toxicity (including mainly standard endpoints such as immobilization or mortality) is also presented in the Supp.info 5 (Table 1). It might be concluded that most of the available data refers to BPA and the information about the toxicity of BPA alternatives is very inconsistent and limited. Even less data is available for the assessment of long-term effects of BPA alternatives towards *D. magna* (Supp.info 5 – Tab 1).

With respect to the prolonged exposure of *D. manga* to BPA and its alternatives, many different endpoints were investigated. This specifically includes not only reproductive indicators (such as: time to first brood, number of first brood per female, brood times, total number of offspring, number of offspring per brood per female and intrinsic rate of natural increase) as well as growth and development parameters (including swimming behavior, body length, heart rate and thoracic limb activity), but also biochemical indicators (acetyl choline esterase (AChE), superoxide dismutase (SOD), catalase (CAT), and glutathione S-transferase (GST) activities as well as malonic dialdehyde (MDA) content) (Liu et al., 2020).

Based on the available results it was confirmed that relative activity
of these enzymes in *D. magna* decreased after exposure to BPA, BPF, BPS
and their mixtures, implying that they may cause the disorder of
digestive, nervous and antioxidative system of *D. magna* (Liu et al.,
2019). Similar observation were also made in the case of BPF, which
induced the disturbance of neurological function and oxidative stress
situation in *D. magna*, leading to behavioral and reproductive effects,
such as growth inhibition, delay in sexual maturation, decreased
spawning, and eventually reduced population fertility (Liu et al., 2020).
Moreover, the observed correlations between biochemical, behavioral,
and reproductive endpoints indicated that the potential of biomarker
techniques as early predictors of bisphenol contaminants-induced al-
terations is also at a higher sensitivity level.

The available data presented in this chapter highlight the fact that
more comprehensive studies on the adverse effects of BPA and its al-
ternatives on aquatic organisms are required at different levels of the
ecological hierarchical scale (Liu et al., 2020). Several marine crusta-
ceans have been proposed as potential candidates to investigate the
biological effects of contaminants on primary consumers. Brine shrimp
*Artemia* spp. (Crustacea, Branchiopoda, Anostraca), commonly known
as sea monkey, is a cosmopolitan organism inhabiting saline environ-
ments, such as inland saltwater lakes and coastal lagoons, feeding pri-
marily on phytoplankton and being an important primary consumer.
The main advantages of brine shrimp which make it a good test or-
ganism is the short life cycle thus a short test duration (i.e. 28–72 h from
hatching to the first endpoint) and the easy culturing, maintenance and
manipulation under laboratory conditions. Moreover, the use of nauplii
hatched from commercially available durable cysts (eggs) ensures
population homogeneity (organisms of similar age, genotype and
physiological condition) and availability all year-round without the
necessity of culturing (Nunes et al., 2006).

Artemia's sensitivity, like in most aquatic organisms, depends on its
developmental stage. According to literature, instar II-III (larval stage)
have been shown to be the most sensitive stage (Sorgeloos et al., 1978;
Vanhaecke and Persoone, 1984). One of the most important variables in
the performance of any biological experiment is the nutrition of the test
animals (Bengtson et al., 1984). In this context, one of the important
advantages of Artemia is that nauplii (stage II-III) do not require feeding
for up to 48 h. Therefore, in assessing the ecotoxicity of pollutants, the
acute toxicity test with mortality/immobilization (after 5–30 s obser-
vation) as the end point is most often used. Besides, sublethal indicators
of toxicity have already been applied for Artemia as biological endpoints,
such as biomarkers (acetylcholinesterase; heat shock proteins; lipid
peroxidation; enzymatic biomarkers of oxidative stress such as glutathione peroxidase; glutathione S-transferase; glutathione reductase), hatching (dry biomass, morphological disorders and size), behavioral changes (swimming speed and path length), and genotoxicity. Long-term chronic tests focus on growth, reproduction and survival or mortality after 7–28 d exposure from larval to adult stage (Libralato et al., 2016).

Very good interspecies correlation between *A. salina* and other important aquatic organisms used for ecotoxicological testing has been observed. It was concluded that *A. salina* could be used as an appreciable tool for prediction of toxicity to other organisms, especially to fish (Kalcíková et al., 2012). *Artemia* spp. was also used for the acute toxicity assessment of BPA (Table 5). Unfortunately, ecotoxicity data for BPA alternatives obtained using this species are not available so far.

The results of a study of influence of brominated bisphenol TBBPA on survival (LC₅₀) of the brine shrimp *A. salina* indicated that brine shrimp is more susceptible to TBBPA toxicity in comparison to other aquatic organisms such as microalgae, crustaceans (*D. magna*, copepods) and fish. The study revealed that *Artemia* can also be used in tests based on biomarkers, such as enzymatic activity. The impact of sublethal concentrations (LC₂₅) of TBPA on the activity of enzymes involved in the cellular antioxidant system (GST and GPx), energy metabolism (LDH) and neurotoxicity marker (AChE) was also observed (Tapia-Salazar et al., 2022).

<table><caption>**Table 5**
The acute toxicity of BPA to Artemia species.</caption><thead><tr><th>Species life stage</th><th>Endpoint</th><th>LC<sub>50</sub> [mg/L]</th><th>Other observations/conclusions</th><th>Reference</th></tr></thead><tbody><tr><td>Artemia franciscana nauplii (instar II-III)</td><td>mortality</td><td>LC<sub>50, 24h</sub> = 44.8; LC<sub>50, 48h</sub> = 34.7</td><td>BPA exposure influenced nauplii growth; dose-length relationship was observed</td><td>Castritsi-Catharios et al. (2013)</td></tr><tr><td>Artemia franciscana nauplii (instar II-III)</td><td>mortality</td><td>LC<sub>50,24h</sub> = 45.51; LC<sub>50, 48h</sub> = 34.45, LC<sub>50,72h</sub> = 17.12</td><td>growth inhibition was observed and was suggested as a valid endpoint for toxicity studies</td><td>Ekonomou et al. (2019)</td></tr><tr><td>Artemia salina adult</td><td>mortality</td><td>LC<sub>50,96h</sub> = 107.2</td><td>BPA toxicity increased towards higher trophic levels (microalga (Tetraselmis sp.) &lt; zooplanktonic grazer (A. salina) &lt; deposit-feeder invertebrate (Heleobia australis) &lt; omnivorous fish (Poecilia vivipara)</td><td>Naveira et al. (2021)</td></tr><tr><td>Artemia salina nauplii (instar II-III)</td><td>immobilization</td><td>EC<sub>50, 24h</sub> = 56.1</td><td></td><td>Kalcíková et al. (2012)</td></tr><tr><td>Artemia sinica</td><td>mortality</td><td>LC<sub>50, 24h</sub> = 70.1; LC<sub>50, 48h</sub> = 50.4; LC<sub>50, 72h</sub> = 17.3</td><td></td><td>Shaukat et al. (2014)</td></tr><tr><td>Artemia sp. nauplii (instar II)</td><td>mortality</td><td>LC<sub>50, 24h</sub> = 74.7; LC<sub>50, 48h</sub> = 59.4</td><td>PNEC (24 h) = 74.6; (48) = 59.3; RQ (24 h) = 1.02 × 10<sup>-6</sup> (48 h) = 1.29 × 10<sup>-6</sup></td><td>da Silva and de Souza Abessa (2019)</td></tr></tbody></table>

In all studies listed in Supp.info 5 (Table 1), time and concentration-dependent effect of BPA on mortality was observed. The other effect of BPA which is inhibition of *Artemia* nauplii growth can be estimated within a short exposure period (24 h), even at doses lower than the median lethal concentration (Castritsi-Catharios et al., 2013). It was suggested that mechanisms of BPA action in *Artemia* may be related with its endocrine disrupting activity. This is in line with the effects summarised in insects' section, since BPA has been demonstrated as a weak ecdysteroid antagonist in insects (Dinan et al., 2001) and in daphnids (Mu et al., 2005). In arthropods, including crustaceans, ecdysteroid hormones participate as endocrine-signaling molecules and control biological functions like molting, growth, reproduction and embryogenesis (Song et al., 2017; LaFont, 2000; Lafont and Mathieu, 2007).

4.5.2.5. *Nematoda. Caenorhabditis elegans* is a well-established model organism with many favorable characteristics for studying toxic effects at different developmental stages and across several generations. *C. elegans* are relatively easy and cheap to maintain, have a short life cycle of approximately four days (at 20 °C) and are highly reproductive. These are prime assets for reproductive and transgenerational studies and thus, the assessment of endocrine disruptive effects (Höss and Weltje, 2007). *C. elegans* is a hermaphroditic nematode species with bilobed gonads which, under certain circumstances (like stress or toxic exposure), develop a small minority (~0.2 %) of XO males by non-disjunction of the X chromosome (Hunt, 2017), which produce sperm and mate with the hermaphrodites. The transparency of the worms allows for easy insight into the germ cell development and the processes during meiotic differentiation, spermiation, ovulation and fertilization. This offers the opportunity to study these processes under chemical exposure to identify germ cell toxicity. Also, the manifold of available mutant and fluorescent reporter strains of *C. elegans* facilitates mechanical studies *in-vivo* in the live worms.

In terms of BPA and its alternatives, many studies have already demonstrated that *C. elegans* is a suitable model to study the adverse effects of these compounds, from reproductive toxicity to neurotoxicity, which are known or suspected from mammalian and non-mammalian vertebrate studies. There is good evidence that findings obtained from *C. elegans* studies translate into effects in higher organisms (Kaletta and Hengartner, 2006; Leung et al., 2008). The less complex biology combined with the high level of genetic and functional conservation and characterization certainly favor the development of *C. elegans* based NAMs. Invertebrate whole organism models like *C. elegans* are legally unprotected animals (in Europe, except for decapodes and cephalopods) and therefore represent animal-alternatives, and they fulfill the requirements for medium to high-throughput approaches (reviewed by Cornaglia et al. 2017). Moreover, *C. elegans* is equally applicable in the field of toxicology as well as ecotoxicology. Particularly attractive for the ecotoxicology field is *C. elegans*' amenability to artificial and diverse experimental environments and conditions. Toxicity testing can be performed with different media, from agar to soil, sediment and to aqueous media and petri dishes work equally well to microwell-plates. Above all, *C. elegans* represents the ecologically important and biologically diverse organism group of nematodes, which inhabit and often dominate most terrestrial and aquatic habitats (reviewed in Yeates et al. 2009, (Yeates et al., 2009). Nematodes can thus act as important indicators of ecosystem health (Höss et al., 2006).

However, there are some limitations that have to be taken into consideration when using *C. elegans* for toxicological studies. *C. elegans* are unable to synthesise cholesterol and therefore require food sources containing sufficient cholesterol or supplementation. Depletion leads to impairment of development, molting, growth and reproduction (Merris et al., 2004; Entchev and Kurzchalia, 2005). BPA and structurally similar BPA-analogues like BPS, impact steroidogenesis and it has been shown that this effect is mediated by the disruption of the cholesterol transport between the outer and the inner mitochondrial membrane (Chen et al.,
peroxidation; enzymatic biomarkers of oxidative stress such as glutathione peroxidase; glutathione S-transferase; glutathione reductase), hatching (dry biomass, morphological disorders and size), behavioral changes (swimming speed and path length), and genotoxicity. Long-term chronic tests focus on growth, reproduction and survival or mortality after 7–28 d exposure from larval to adult stage (Libralato et al., 2016).

Very good interspecies correlation between *A. salina* and other important aquatic organisms used for ecotoxicological testing has been observed. It was concluded that *A. salina* could be used as an appreciable tool for prediction of toxicity to other organisms, especially to fish (Kalcíková et al., 2012). *Artemia* spp. was also used for the acute toxicity assessment of BPA (Table 5). Unfortunately, ecotoxicity data for BPA alternatives obtained using this species are not available so far.

The results of a study of influence of brominated bisphenol TBBPA on survival (LC₅₀) of the brine shrimp *A. salina* indicated that brine shrimp is more susceptible to TBBPA toxicity in comparison to other aquatic organisms such as microalgae, crustaceans (*D. magna*, copepods) and fish. The study revealed that *Artemia* can also be used in tests based on biomarkers, such as enzymatic activity. The impact of sublethal concentrations (LC₂₅) of TBPA on the activity of enzymes involved in the cellular antioxidant system (GST and GPx), energy metabolism (LDH) and neurotoxicity marker (AChE) was also observed (Tapia-Salazar et al., 2022).

<table><caption>**Table 5**
The acute toxicity of BPA to Artemia species.</caption><thead><tr><th>Species life stage</th><th>Endpoint</th><th>LC<sub>50</sub> [mg/L]</th><th>Other observations/conclusions</th><th>Reference</th></tr></thead><tbody><tr><td>Artemia franciscana nauplii (instar II-III)</td><td>mortality</td><td>LC<sub>50</sub>, 24h = 44.8; LC<sub>50</sub>, 48h = 34.7</td><td>BPA exposure influenced nauplii growth; dose-length relationship was observed</td><td>Castritsi-Catharios et al. (2013)</td></tr><tr><td>Artemia franciscana nauplii (instar II-III)</td><td>mortality</td><td>LC<sub>50</sub>, 24h = 45.51; LC<sub>50</sub>, 48h = 34.45, LC<sub>50</sub>, 72h = 17.12</td><td>growth inhibition was observed and was suggested as a valid endpoint for toxicity studies</td><td>Ekonomou et al. (2019)</td></tr><tr><td>Artemia salina adult</td><td>mortality</td><td>LC<sub>50</sub>, 96h = 107.2</td><td>BPA toxicity increased towards higher trophic levels (microalga (Tetraselmis sp.) &lt; zooplanktonic grazer (A. salina) &lt; deposit-feeder invertebrate (Heleobia australis) &lt; omnivorous fish (Poecilia vivipara)</td><td>Naveira et al. (2021)</td></tr><tr><td>Artemia salina nauplii (instar II-III)</td><td>immobilization</td><td>EC<sub>50</sub>, 24h = 56.1</td><td></td><td>Kalcíková et al. (2012)</td></tr><tr><td>Artemia sinica</td><td>mortality</td><td>LC<sub>50</sub>, 24h = 70.1; LC<sub>50</sub>, 48h = 50.4; LC<sub>50</sub>, 72h = 17.3</td><td></td><td>Shaukat et al. (2014)</td></tr><tr><td>Artemia sp. nauplii (instar II)</td><td>mortality</td><td>LC<sub>50</sub>, 24h = 74.7; LC<sub>50</sub>, 48h = 59.4</td><td>PNEC (24 h) = 74.6; (48) = 59.3; RQ (24 h) = 1.02 × 10<sup>-6</sup> (48 h) = 1.29 × 10<sup>-6</sup></td><td>da Silva and de Souza Abessa (2019)</td></tr></tbody></table>

In all studies listed in Supp.info 5 (Table 1), time and concentration-dependent effect of BPA on mortality was observed. The other effect of BPA which is inhibition of *Artemia* nauplii growth can be estimated within a short exposure period (24 h), even at doses lower than the median lethal concentration (Castritsi-Catharios et al., 2013). It was suggested that mechanisms of BPA action in *Artemia* may be related with its endocrine disrupting activity. This is in line with the effects summarised in insects' section, since BPA has been demonstrated as a weak ecdysteroid antagonist in insects (Dinan et al., 2001) and in daphnids (Mu et al., 2005). In arthropods, including crustaceans, ecdysteroid hormones participate as endocrine-signaling molecules and control biological functions like molting, growth, reproduction and embryogenesis (Song et al., 2017; LaFont, 2000; Lafont and Mathieu, 2007).

4.5.2.5. *Nematoda. Caenorhabditis elegans* is a well-established model organism with many favorable characteristics for studying toxic effects at different developmental stages and across several generations. *C. elegans* are relatively easy and cheap to maintain, have a short life cycle of approximately four days (at 20 °C) and are highly reproductive. These are prime assets for reproductive and transgenerational studies and thus, the assessment of endocrine disruptive effects (Höss and Weltje, 2007). *C. elegans* is a hermaphroditic nematode species with bilobed gonads which, under certain circumstances (like stress or toxic exposure), develop a small minority (~0.2 %) of XO males by non-disjunction of the X chromosome (Hunt, 2017), which produce sperm and mate with the hermaphrodites. The transparency of the worms allows for easy insight into the germ cell development and the processes during meiotic differentiation, spermiation, ovulation and fertilization. This offers the opportunity to study these processes under chemical exposure to identify germ cell toxicity. Also, the manifold of available mutant and fluorescent reporter strains of *C. elegans* facilitates mechanical studies *in-vivo* in the live worms.

In terms of BPA and its alternatives, many studies have already demonstrated that *C. elegans* is a suitable model to study the adverse effects of these compounds, from reproductive toxicity to neurotoxicity, which are known or suspected from mammalian and non-mammalian vertebrate studies. There is good evidence that findings obtained from *C. elegans* studies translate into effects in higher organisms (Kaletta and Hengartner, 2006; Leung et al., 2008). The less complex biology combined with the high level of genetic and functional conservation and characterization certainly favor the development of *C. elegans* based NAMs. Invertebrate whole organism models like *C. elegans* are legally unprotected animals (in Europe, except for decapodes and cephalopods) and therefore represent animal-alternatives, and they fulfill the requirements for medium to high-throughput approaches (reviewed by Cornaglia et al. 2017). Moreover, *C. elegans* is equally applicable in the field of toxicology as well as ecotoxicology. Particularly attractive for the ecotoxicology field is *C. elegans*' amenability to artificial and diverse experimental environments and conditions. Toxicity testing can be performed with different media, from agar to soil, sediment and to aqueous media and petri dishes work equally well to microwell-plates. Above all, *C. elegans* represents the ecologically important and biologically diverse organism group of nematodes, which inhabit and often dominate most terrestrial and aquatic habitats (reviewed in Yeates et al. 2009, (Yeates et al., 2009). Nematodes can thus act as important indicators of ecosystem health (Höss et al., 2006).

However, there are some limitations that have to be taken into consideration when using *C. elegans* for toxicological studies. *C. elegans* are unable to synthesise cholesterol and therefore require food sources containing sufficient cholesterol or supplementation. Depletion leads to impairment of development, molting, growth and reproduction (Merris et al., 2004; Entchev and Kurzchalia, 2005). BPA and structurally similar BPA-analogues like BPS, impact steroidogenesis and it has been shown that this effect is mediated by the disruption of the cholesterol transport between the outer and the inner mitochondrial membrane (Chen et al.,
2019). It has been shown that cholesterol supplementation under experimental laboratory conditions has the ability to attenuate reproductive effects in *C. elegans* (Chen et al., 2019; Chen et al., 2016; Gibbs, 2018). The majority of *C. elegans* studies in the toxicological context add *E. coli* as food supply to their test system. For toxicological examinations, the introduction of another organism with metabolic capacities represents a substantial confounding factor in an exposure scenario (Hunt, 2017). The associated issue of xenobiotic metabolism is not the only aspect; the capacity of the bacteria to act as vehicles for the toxicants as well as recipient adds another level of complexity to the exposure and the toxicokinetics.

Among bisphenols, BPA is probably the best studied compound in *C. elegans*, followed by BPS, BPF, BPAP, TBBPA; TBBPF and TCBPA. The number of publications for each of these compounds according to our research at the time of this publication, was > 20; one publication explored the effects of BPY and BPZ in *C. elegans* (Roberts et al., 2022). The most commonly assessed endpoints in all of these studies comprised growth, reproduction, behaviors (locomotion, body bends, head thrashes, mechano-sensory habituation, pharynx pumping rate or defecation), DNA-damage (e.g., measurement of DNA damage checkpoint kinase CHK-1 activation), germ cell apoptosis (e.g., acridine orange staining), stress-related gene expression (e.g. of *hsp-3*, *hsp-4*, *hsp-16.1*, *16.2*, and *hsp-70*) oxidative stress (detection of ROS and anti-oxidative enzymes like superoxide dismutase (SOD), catalase (CAT), glutathione-S-transferase (GST), at protein and mRNA-level) (Tan et al., 2015; García-Espíneira et al., 2018). All above mentioned endpoints were found to be affected by exposure to BPA but at varying effect threshold levels between 0.1 and 1.0 µM.

BPA was also able to induce the expression of the vitellogenin gene *vit-1* in *C. elegans* in a concentration-dependent manner (Cicirra and David, 2020). The gene *vit-1* is expressed in the intestine of the worms and the transcription factor *mab-3* (male-abnormal gene) represses *vit-1* in male *C. elegans*. The study provides evidence that *mab-3* is repressed by BPA and that this may be linked to the increase in *vit-1* expression.

Epidemiological as well as animal studies have shown that BPA and BPS can have obesogenic effects through a variety of mechanisms like disruption of glucose homeostasis and lipogenesis, inflammation, and oxidative stress (Wei et al., 2011; Xiao et al., 2021; Rezg et al., 2014) was able to reproduce these obesogenic effects in *C. elegans*, demonstrating that exposure to BPS increases fat accumulation in the nematodes and that this effect is most likely mediated by an upregulation of fatty acids biosynthesis (via upregulation of the gene *daf-16* in the insulin signaling pathway of *C. elegans*) and inhibition of fatty acid oxidation. Down-regulation of the nuclear receptor gene *nhr-49*, which has been identified as one of the key regulators of fat metabolism in *C. elegans* (Van Gilst et al., 2005), appears to play a central role in the BPS elicited obesogenic effects (Xiao et al., 2021; Zhou et al., 2021). Similarly, a concentration-dependent increase in lipid deposition was also found in worms exposed to BPA (García-Espíneira et al., 2018).

Lipophilic hormone signaling between the reproductive system and the intestine of the worms can influence adult longevity (Berman and Kenyon, 2006) via the induction of the gene *kri-1* which signals the nuclear translocation and activation of the DAF-16 transcription factor. However, this only happens in germline deficient worms, hence indicating that the aging process is related to the maturation and proliferation of germ cells. Bisphenol compounds have shown to induce oxidative stress in germline cells of *C. elegans* and that this can lead to DNA-damage and subsequently to germ cell apoptosis and chromosome abnormalities (Allard and Colaiacovo, 2010; Hornos Carneiro et al., 2020). These effects of bisphenols do not only manifest as reproductive toxicity (but also promotes degenerative age-related responses, accelerating the aging process in *C. elegans* (Tan et al., 2015).

*C. elegans* appears as a promising organism for the development of NAMs to be used in the context of an invertebrate EDC testing framework. Moreover, chemical screening approaches with *C. elegans* in combination with computational models, hold great potential for the

identification of MIEs and KE and thus, the development of invertebrate-specific AOPs.

4.5.2.6. Annelida (earthworms). Earthworms are one of the most widely used soil organisms in ecotoxicological studies, as they are excellent bioindicators of contamination (Marcos et al., 2021) and have different standardized tests for toxicity assessment (OECD, 2016, 1984) (Test No. 222, 2016; Test No. 207, 1984). These organisms are vital for maintaining food chains (Novo et al., 2018), so their ability to incorporate and accumulate soil contaminants orally or by passive diffusion through the skin can ultimately negatively affect populations of other species. In addition, earthworms may affect soil degradation and the transformation of organic pollutants (Qian et al., 2023). Their hermaphroditism makes them excellent ecological candidates for assessing the effects of endocrine disruptors on the male and female reproductive systems in the same organism (Il Kwak and An, 2021).

Several annelid species have been used to assess the toxic effects of BPA or its alternatives. Since the sensitivity of earthworms to the same chemicals differs even between closely related species, it is helpful to consider results obtained with different species for a more robust toxicological assessment (Marcos et al., 2021).

Two earthworm species (*Dendrobaena veneta* and *Eisenia fetida*) were tested against BPS and compared to BPA toxicity aiming at suggesting it as a potential less toxic alternative (Marcos et al., 2021). The two species showed different patterns of response towards BPS at 21 °C. In the case of *E. fetida*, no mortality was observed up to the highest concentration tested (1000 mg/kg artificial soil), with growth being stimulated in the first 14 days but then just showing a trend of decrease in weight similar to the control. In the reproduction test, the number of juveniles was lower at the highest concentration than in the control. In the case of *D. veneta*, no effects were observed regarding their survival, growth or reproduction. In the same study similar experiments were also performed with *D. veneta* at 26 °C, and some changes were observed, with 10 mg BPS/kg inducing a weight increase higher than the control in the first 14 days of exposure. However, this pattern changed in the last 14 days of exposure, with weight gain observed at 100 mg BPS/kg, while the lowest (10 mg BPS/kg) and the highest (1000 mg BPS/kg) concentrations tested showed earthworms decrease in weight. In a contact test, where only dermal exposure was tested, BPS induced mortality after 48 h at 2 mg/ml to these two earthworm species. In addition, earthworms showed oxidative stress responses at 21 °C, with GST activity decreasing for *E. fetidawhile increasing for D. veneta*. In a separate study (Verdú et al., 2018), these two species were subjected to a contact test, and reproduction traits were only significantly different for *E. fetida*, for which the number of juveniles decreased at higher concentrations (1000–2000 mg/kg), thus showing different sensitivity in both species.

Conversely, another study with BPA and BPS, this time using *Lumbriculus variegatus*, showed that the effects of both chemicals were nearly indistinguishable after 5 min and 5-day exposures. Both chemicals retarded the initial phase of body regrowth after cutting/fragmentation and increased the pulse rate of the dorsal blood vessel (DBV) at concentrations of 10⁻⁹ or 10⁻⁶ M (Vought and Wang, 2018).

Also, in *L. variegatus*, 1.36 mg/L BPA in the overlying water (57 mg/kg dry weight in sediment) significantly induced biomass loss and a decline in the number of individuals after 28 days of exposure (Staples et al., 2016). BPA concentrations above 130 mg/kg significantly induced the death of *E. fetida* individuals in 7-day and 14-day exposures (Lee et al., 2005), but mortality in this species was also detected in acute 24-h and 48-h exposures to BPA concentrations above 2.42 µg/L (de Oliveira et al., 2021).

The exceptional sensitivity of earthworm species to some estrogens makes it a possible tool for estrogenic EDC screening. In this regard, a test with *L. variegatus* exposed to BPA and BPS demonstrated that both compounds had a similar effect leading to a significant increase in the pulse rate of the DBV at exposures of up to 10 days in the 10⁻¹⁰ to 10⁻⁸
2019). It has been shown that cholesterol supplementation under experimental laboratory conditions has the ability to attenuate reproductive effects in *C. elegans* (Chen et al., 2019; Chen et al., 2016; Gibbs, 2018). The majority of *C. elegans* studies in the toxicological context add *E. coli* as food supply to their test system. For toxicological examinations, the introduction of another organism with metabolic capacities represents a substantial confounding factor in an exposure scenario (Hunt, 2017). The associated issue of xenobiotic metabolism is not the only aspect; the capacity of the bacteria to act as vehicles for the toxicants as well as recipient adds another level of complexity to the exposure and the toxicokinetics.

Among bisphenols, BPA is probably the best studied compound in *C. elegans*, followed by BPS, BPF, BPAP, TBBPA; TBBPF and TCBPA. The number of publications for each of these compounds according to our research at the time of this publication, was > 20; one publication explored the effects of BPY and BPZ in *C. elegans* (Roberts et al., 2022). The most commonly assessed endpoints in all of these studies comprised growth, reproduction, behaviors (locomotion, body bends, head thrashes, mechano-sensory habituation, pharynx pumping rate or defecation), DNA-damage (e.g., measurement of DNA damage checkpoint kinase CHK-1 activation), germ cell apoptosis (e.g., acridine orange staining), stress-related gene expression (e.g. of *hsp-3*, *hsp-4*, *hsp-16.1*, *16.2*, and *hsp-70*) oxidative stress (detection of ROS and anti-oxidative enzymes like superoxide dismutase (SOD), catalase (CAT), glutathione-S-transferase (GST), at protein and mRNA-level) (Tan et al., 2015; García-Espíneira et al., 2018). All above mentioned endpoints were found to be affected by exposure to BPA but at varying effect threshold levels between 0.1 and 1.0 µM.

BPA was also able to induce the expression of the vitellogenin gene *vit-1* in *C. elegans* in a concentration-dependent manner (Cicirra and David, 2020). The gene *vit-1* is expressed in the intestine of the worms and the transcription factor *mab-3* (male-abnormal gene) represses *vit-1* in male *C. elegans*. The study provides evidence that *mab-3* is repressed by BPA and that this may be linked to the increase in *vit-1* expression.

Epidemiological as well as animal studies have shown that BPA and BPS can have obesogenic effects through a variety of mechanisms like disruption of glucose homeostasis and lipogenesis, inflammation, and oxidative stress (Wei et al., 2011; Xiao et al., 2021; Rezg et al., 2014) was able to reproduce these obesogenic effects in *C. elegans*, demonstrating that exposure to BPS increases fat accumulation in the nematodes and that this effect is most likely mediated by an upregulation of fatty acids biosynthesis (via upregulation of the gene *daf-16* in the insulin signaling pathway of *C. elegans*) and inhibition of fatty acid oxidation. Down-regulation of the nuclear receptor gene *nhr-49*, which has been identified as one of the key regulators of fat metabolism in *C. elegans* (Van Gilst et al., 2005), appears to play a central role in the BPS elicited obesogenic effects (Xiao et al., 2021; Zhou et al., 2021). Similarly, a concentration-dependent increase in lipid deposition was also found in worms exposed to BPA (García-Espíneira et al., 2018).

Lipophilic hormone signaling between the reproductive system and the intestine of the worms can influence adult longevity (Berman and Kenyon, 2006) via the induction of the gene *kri-1* which signals the nuclear translocation and activation of the DAF-16 transcription factor. However, this only happens in germline deficient worms, hence indicating that the aging process is related to the maturation and proliferation of germ cells. Bisphenol compounds have shown to induce oxidative stress in germline cells of *C. elegans* and that this can lead to DNA-damage and subsequently to germ cell apoptosis and chromosome abnormalities (Allard and Colaiacovo, 2010; Hornos Carneiro et al., 2020). These effects of bisphenols do not only manifest as reproductive toxicity (but also promotes degenerative age-related responses, accelerating the aging process in *C. elegans* (Tan et al., 2015).

*C. elegans* appears as a promising organism for the development of NAMs to be used in the context of an invertebrate EDC testing framework. Moreover, chemical screening approaches with *C. elegans* in combination with computational models, hold great potential for the

identification of MIEs and KE and thus, the development of invertebrate-specific AOPs.

4.5.2.6. Annelida (earthworms). Earthworms are one of the most widely used soil organisms in ecotoxicological studies, as they are excellent bioindicators of contamination (Marcos et al., 2021) and have different standardized tests for toxicity assessment (OECD, 2016, 1984) (Test No. 222, 2016; Test No. 207, 1984). These organisms are vital for maintaining food chains (Novo et al., 2018), so their ability to incorporate and accumulate soil contaminants orally or by passive diffusion through the skin can ultimately negatively affect populations of other species. In addition, earthworms may affect soil degradation and the transformation of organic pollutants (Qian et al., 2023). Their hermaphroditism makes them excellent ecological candidates for assessing the effects of endocrine disruptors on the male and female reproductive systems in the same organism (Il Kwak and An, 2021).

Several annelid species have been used to assess the toxic effects of BPA or its alternatives. Since the sensitivity of earthworms to the same chemicals differs even between closely related species, it is helpful to consider results obtained with different species for a more robust toxicological assessment (Marcos et al., 2021).

Two earthworm species (*Dendrobaena veneta* and *Eisenia fetida*) were tested against BPS and compared to BPA toxicity aiming at suggesting it as a potential less toxic alternative (Marcos et al., 2021). The two species showed different patterns of response towards BPS at 21 °C. In the case of *E. fetida*, no mortality was observed up to the highest concentration tested (1000 mg/kg artificial soil), with growth being stimulated in the first 14 days but then just showing a trend of decrease in weight similar to the control. In the reproduction test, the number of juveniles was lower at the highest concentration than in the control. In the case of *D. veneta*, no effects were observed regarding their survival, growth or reproduction. In the same study similar experiments were also performed with *D. veneta* at 26 °C, and some changes were observed, with 10 mg BPS/kg inducing a weight increase higher than the control in the first 14 days of exposure. However, this pattern changed in the last 14 days of exposure, with weight gain observed at 100 mg BPS/kg, while the lowest (10 mg BPS/kg) and the highest (1000 mg BPS/kg) concentrations tested showed earthworms decrease in weight. In a contact test, where only dermal exposure was tested, BPS induced mortality after 48 h at 2 mg/ml to these two earthworm species. In addition, earthworms showed oxidative stress responses at 21 °C, with GST activity decreasing for *E. fetidawhile increasing for D. veneta*. In a separate study (Verdú et al., 2018), these two species were subjected to a contact test, and reproduction traits were only significantly different for *E. fetida*, for which the number of juveniles decreased at higher concentrations (1000–2000 mg/kg), thus showing different sensitivity in both species.

Conversely, another study with BPA and BPS, this time using *Lumbriculus variegatus*, showed that the effects of both chemicals were nearly indistinguishable after 5 min and 5-day exposures. Both chemicals retarded the initial phase of body regrowth after cutting/fragmentation and increased the pulse rate of the dorsal blood vessel (DBV) at concentrations of 10⁻⁹ or 10⁻⁶ M (Vought and Wang, 2018).

Also, in *L. variegatus*, 1.36 mg/L BPA in the overlying water (57 mg/kg dry weight in sediment) significantly induced biomass loss and a decline in the number of individuals after 28 days of exposure (Staples et al., 2016). BPA concentrations above 130 mg/kg significantly induced the death of *E. fetida* individuals in 7-day and 14-day exposures (Lee et al., 2005), but mortality in this species was also detected in acute 24-h and 48-h exposures to BPA concentrations above 2.42 µg/L (de Oliveira et al., 2021).

The exceptional sensitivity of earthworm species to some estrogens makes it a possible tool for estrogenic EDC screening. In this regard, a test with *L. variegatus* exposed to BPA and BPS demonstrated that both compounds had a similar effect leading to a significant increase in the pulse rate of the DBV at exposures of up to 10 days in the 10⁻¹⁰ to 10⁻⁸
M range (Wang and Wang, 2021). BPA has been shown to induce endocrine disruption related to the male reproductive organs of *E. fetida* (Novo et al., 2018), with changes in gene expression of genes EcR, MAPR, AdipoR, along with oxidative stress and multiple histopathological changes in the body wall and ovaries. In a 7-day test with *Eisenia andrei*, BPA inhibited normal oogenesis and maturation of oocytes in earthworm ovary and was observed to damage seminal vesicle tissues and inhibit normal spermatogenesis (Il Kwak and An, 2021).

Behavioral changes have also been detected in *E. fetida* at the individual level, with an increase in mechanical stimulation upon 3-day exposure to BPA 100 nM and 10 µM BPA, together with metabolical (increased lipid oxidation) and physiological (hyperplasia of epidermis, increased body wall thickness and ovarian atrophy) alterations (Babić et al., 2016). On the other hand, also in *E. fetida*, BPS appeared to induce oxidative stress and the process of antioxidant defense in exposures up to 21 days at 1 and 10 mg/kg, altering the superoxide dismutase (SOD) and catalase (CAT) enzymatic activities, and also the expression profiles of the genes annetocin (ANN) and calreticulin (CRT) (Qian et al., 2023).

#### 4.5.2.7. Cnidaria (Hydra sp.).
Hydra sp. are small freshwater dwelling organisms that belong to the phylum Cnidaria, usually inhabit slow running waters and ponds, attached to submerged subtracts (e.g. rocks, macrophytes), and are carnivorous, primarily feeding on small invertebrates. Though Hydra sp. occupy a low trophic level in the food web, acting as predator and prey, they play an important role in structuring the planktonic communities of ponds, being pointed out as a suitable bioindicator (Quinn et al., 2012). In fact, since the eighties, Hydra sp. has been used as model species to evaluate the hazards of environmental contaminants, being recognized as a relevant and sensitive model organism due to several advantageous characteristics (Quinn et al., 2012; Galliot, 2012; Vimalkumar et al., 2022). Being a small organism, easy to culture in the laboratory, and with a fast reproduction and short generation time, it is appropriate for use in small-scale toxicity assays, allowing to easily test several chemicals and concentrations. Hydra sp. are diploblastic organisms (i.e. possess 2 cell layers: the outer ectoderm and the inner endoderm) facilitating the contact of all its cells with the contaminants present in the surrounding medium, contributing to the high sensitivity of these organisms to several types of chemicals. They have a high capacity to regenerate the whole body, which allows assessing teratogenic effects on stem cells. Adding to this endpoint, they can provide comprehensive toxicological analysis integrating several endpoints at individual level (e.g., morphology, feeding, budding, regeneration, reproduction), which constitute a gain when studying the hazards that bisphenols may pose to freshwater biota.

Only a few studies have addressed the ecotoxicity data of bisphenols in *Hydra* sp. species. To our knowledge, four studies have been published addressing the toxicity of BPA and BADGE to threespecies of Hydra: *H. vulgaris*, *H. magnipapillata*, and *H. oligactis*. Pascoe et al. (2002) (*Pascoe et al., 2002*) studied the effects of BPA on *H. vulgaris*, recording that concentrations above 42 µg/L induced morphological changes in the hydras, and calculating a LC₅₀,₉₶h of 6.9 mg/L for this endpoint. Significant effects were also observed in the regeneration of hydra at 4.6 mg/L of BPA: sections of hydra could only regenerate tentacle buds. Later, Murgardas et al (2019) (Murugadas et al., 2019) reported that *H. magnipapillata* exposed to BPA for 96 h, also went through notorious morphological changes, calculating a slightly lower LC₅₀,₉₶h than the one reported by Pascoe et al. (2002) (*Pascoe et al., 2002*): 5.137 mg/L. These authors also observed significant delays on head regeneration in hydras exposed to 15 µM (3.42 mg/L), hypothesizing that these effects were due to the cytotoxicity caused by BPA and the production of ROS at the amputated region which could promote DNA damage and cell death. Polyps exposed to this same BPA concentration also exhibited difficulties in capturing the prey, which could be related with failure in nematocysts discharge to paralyze the prey and on opening the mouth. The effect of BPA on the regeneration of gastric sections of this same

species of hydra was studied by Park and Yeo (2012) (Park and Yeo, 2012). Hydra sections exposed up to 5 ppm (5 mg/L) exhibited a regeneration process slower than those exposed in the control. Observations at the microscopic level revealed no injuries in the nematocysts, but cell apoptosis occurred in hydra exposed to BPA. After transferring the BPA-exposed hydras to the control medium, the authors observed that they recovered their regenerative capacity, and that the third generation showed a reproductive capacity similar to the ones from the control group. The effects of BPA, at similar concentrations as those tested by Park and Yeo (2012) (Park and Yeo, 2012), on the reproduction of *H. oligactis*, was demonstrated by Fukuhori et al. (2005) (Fukuhori et al., 2005). Males and females of *H. oligactis* with sexual reproduction were exposed to concentrations of BPA ranging from 0.5 to 4 mg/L. In males, testis formation was suppressed at concentrations equal or above 1 mg/L, while tentacles and body length were reduced in organism exposed to 3 mg/L. As for females, the mean number of produced eggs was reduced at concentrations above 0.5 mg/L. Further 1 mg/L stimulated asexual reproduction, but budding production was suppressed at concentrations equal or above 2 mg/L. These authors also studied the accumulation of BPA in the body of hydras, having registered an influence of temperature. BPA concentration in the body of polyps was approximately 1.4 times higher at 10 °C than at 20 °C.

The toxicity of BADGE to *H. magnipapillata* was also reported by Park and Yeo (2012) (Park and Yeo, 2012). Regarding the capacity for regeneration, BADGE induced similar effects as those observed for BPA and at the same concentrations. A reduction in the regeneration process of the hydra sections was observed, followed by a restoration of the regenerative capacity after the hydras being transferred to the control medium. Microscopic observations showed that exposure to BADGE caused apoptosis and cellular necrosis.

#### 4.5.3. Aquatic vertebrates

Our investigation identified 17 BPA alternatives that were toxicologically evaluated on aquatic vertebrate models, specifically employing fish models, namely zebrafish (*Danio rerio*), Mozambique tilapia (*Oreochromis mossambicus*), African catfish (*Clarias gariepinus*), rainbow trout (*Oncorhynchus mykiss*), marine medaka (*Oryzias melastigma*), olive flounder (*Paralichthys olivaceus*), red common carp (*Cyprinus carpio*), rohu fish (*Labeo rohita*), European flounder (*Platichthys flesus*), goldfish (*Carassius auratus*), spotted snakehead (*Channa punctatus*), and brown trout (*Salmo trutta*) and amphibian models, namely Argentine toad (*Rhinella arenarum*), black-spotted pond frog (*Pelophylax nigromaculatus*; formerly *Rana nigromaculata*), Japanese wrinkled frog (*Rana rugosa*), African clawed frog (*Xenopus laevis*), and western clawed frog (*Xenopus tropicalis*; formerly *Silurana tropicalis*); Supp. info. 6.

While many different vertebrate species were studied, zebrafish is by far the most comprehensively studied aquatic vertebrate model organism in terms of the number of tested BPA alternatives as well as assessed endpoints and developmental stages (Supp. info. 6). Importantly, the collected studies show that BPA alternatives have not been systematically studied, i.e., studies that evaluated specific endpoints (e.g., effects related to endocrine disruption) across the models (fish vs. amphibians), or multiple chemicals, are largely lacking. The different chemicals have also been unequally investigated so far, with nearly 70 % of identified studies focusing on the top four BPA alternatives (BPS, BPF, BPAF, and TBBPA) (Fig. 5). Among the chemicals that have been examined, BPS has undergone the most comprehensive research in terms of both the quantity of studies conducted and the range of endpoints explored.

#### 4.5.3.1. Amphibian models.
Compared to fishes, much less studies on BPA and its alternatives are available in amphibians (Supp. info 6). The summary of their effects is visualized in Supp. info. 7. They are restricted to the order Anura, whereas no experiments using other amphibian orders, Urodeles and Caecilians, have been reported. The available studies focused mainly on investigations of BPA effects on sexual differentiation
M range (Wang and Wang, 2021). BPA has been shown to induce endocrine disruption related to the male reproductive organs of *E. fetida* (Novo et al., 2018), with changes in gene expression of genes EcR, MAPR, AdipoR, along with oxidative stress and multiple histopathological changes in the body wall and ovaries. In a 7-day test with *Eisenia andrei*, BPA inhibited normal oogenesis and maturation of oocytes in earthworm ovary and was observed to damage seminal vesicle tissues and inhibit normal spermatogenesis (Il Kwak and An, 2021).

Behavioral changes have also been detected in *E. fetida* at the individual level, with an increase in mechanical stimulation upon 3-day exposure to BPA 100 nM and 10 µM BPA, together with metabolical (increased lipid oxidation) and physiological (hyperplasia of epidermis, increased body wall thickness and ovarian atrophy) alterations (Babić et al., 2016). On the other hand, also in *E. fetida*, BPS appeared to induce oxidative stress and the process of antioxidant defense in exposures up to 21 days at 1 and 10 mg/kg, altering the superoxide dismutase (SOD) and catalase (CAT) enzymatic activities, and also the expression profiles of the genes annetocin (ANN) and calreticulin (CRT) (Qian et al., 2023).

#### 4.5.2.7. Cnidaria (Hydra sp.).
Hydra sp. are small freshwater dwelling organisms that belong to the phylum Cnidaria, usually inhabit slow running waters and ponds, attached to submerged subtracts (e.g. rocks, macrophytes), and are carnivorous, primarily feeding on small invertebrates. Though Hydra sp. occupy a low trophic level in the food web, acting as predator and prey, they play an important role in structuring the planktonic communities of ponds, being pointed out as a suitable bioindicator (Quinn et al., 2012). In fact, since the eighties, Hydra sp. has been used as model species to evaluate the hazards of environmental contaminants, being recognized as a relevant and sensitive model organism due to several advantageous characteristics (Quinn et al., 2012; Galliot, 2012; Vimalkumar et al., 2022). Being a small organism, easy to culture in the laboratory, and with a fast reproduction and short generation time, it is appropriate for use in small-scale toxicity assays, allowing to easily test several chemicals and concentrations. Hydra sp. are diploblastic organisms (i.e. possess 2 cell layers: the outer ectoderm and the inner endoderm) facilitating the contact of all its cells with the contaminants present in the surrounding medium, contributing to the high sensitivity of these organisms to several types of chemicals. They have a high capacity to regenerate the whole body, which allows assessing teratogenic effects on stem cells. Adding to this endpoint, they can provide comprehensive toxicological analysis integrating several endpoints at individual level (e.g., morphology, feeding, budding, regeneration, reproduction), which constitute a gain when studying the hazards that bisphenols may pose to freshwater biota.

Only a few studies have addressed the ecotoxicity data of bisphenols in *Hydra* sp. species. To our knowledge, four studies have been published addressing the toxicity of BPA and BADGE to threespecies of Hydra: *H. vulgaris*, *H. magnipapillata*, and *H. oligactis*. Pascoe et al. (2002) (*Pascoe et al., 2002*) studied the effects of BPA on *H. vulgaris*, recording that concentrations above 42 µg/L induced morphological changes in the hydras, and calculating a LC₅₀,₉₶h of 6.9 mg/L for this endpoint. Significant effects were also observed in the regeneration of hydra at 4.6 mg/L of BPA: sections of hydra could only regenerate tentacle buds. Later, Murgardas et al (2019) (Murugadas et al., 2019) reported that *H. magnipapillata* exposed to BPA for 96 h, also went through notorious morphological changes, calculating a slightly lower LC₅₀,₉₶h than the one reported by Pascoe et al. (2002) (*Pascoe et al., 2002*): 5.137 mg/L. These authors also observed significant delays on head regeneration in hydras exposed to 15 µM (3.42 mg/L), hypothesizing that these effects were due to the cytotoxicity caused by BPA and the production of ROS at the amputated region which could promote DNA damage and cell death. Polyps exposed to this same BPA concentration also exhibited difficulties in capturing the prey, which could be related with failure in nematocysts discharge to paralyze the prey and on opening the mouth. The effect of BPA on the regeneration of gastric sections of this same

species of hydra was studied by Park and Yeo (2012) (Park and Yeo, 2012). Hydra sections exposed up to 5 ppm (5 mg/L) exhibited a regeneration process slower than those exposed in the control. Observations at the microscopic level revealed no injuries in the nematocysts, but cell apoptosis occurred in hydra exposed to BPA. After transferring the BPA-exposed hydras to the control medium, the authors observed that they recovered their regenerative capacity, and that the third generation showed a reproductive capacity similar to the ones from the control group. The effects of BPA, at similar concentrations as those tested by Park and Yeo (2012) (Park and Yeo, 2012), on the reproduction of *H. oligactis*, was demonstrated by Fukuhori et al. (2005) (Fukuhori et al., 2005). Males and females of *H. oligactis* with sexual reproduction were exposed to concentrations of BPA ranging from 0.5 to 4 mg/L. In males, testis formation was suppressed at concentrations equal or above 1 mg/L, while tentacles and body length were reduced in organism exposed to 3 mg/L. As for females, the mean number of produced eggs was reduced at concentrations above 0.5 mg/L. Further 1 mg/L stimulated asexual reproduction, but budding production was suppressed at concentrations equal or above 2 mg/L. These authors also studied the accumulation of BPA in the body of hydras, having registered an influence of temperature. BPA concentration in the body of polyps was approximately 1.4 times higher at 10 °C than at 20 °C.

The toxicity of BADGE to *H. magnipapillata* was also reported by Park and Yeo (2012) (Park and Yeo, 2012). Regarding the capacity for regeneration, BADGE induced similar effects as those observed for BPA and at the same concentrations. A reduction in the regeneration process of the hydra sections was observed, followed by a restoration of the regenerative capacity after the hydras being transferred to the control medium. Microscopic observations showed that exposure to BADGE caused apoptosis and cellular necrosis.

#### 4.5.3. Aquatic vertebrates

Our investigation identified 17 BPA alternatives that were toxicologically evaluated on aquatic vertebrate models, specifically employing fish models, namely zebrafish (*Danio rerio*), Mozambique tilapia (*Oreochromis mossambicus*), African catfish (*Clarias gariepinus*), rainbow trout (*Oncorhynchus mykiss*), marine medaka (*Oryzias melastigma*), olive flounder (*Paralichthys olivaceus*), red common carp (*Cyprinus carpio*), rohu fish (*Labeo rohita*), European flounder (*Platichthys flesus*), goldfish (*Carassius auratus*), spotted snakehead (*Channa punctatus*), and brown trout (*Salmo trutta*) and amphibian models, namely Argentine toad (*Rhinella arenarum*), black-spotted pond frog (*Pelophylax nigromaculatus*; formerly *Rana nigromaculata*), Japanese wrinkled frog (*Rana rugosa*), African clawed frog (*Xenopus laevis*), and western clawed frog (*Xenopus tropicalis*; formerly *Silurana tropicalis*); Supp. info. 6.

While many different vertebrate species were studied, zebrafish is by far the most comprehensively studied aquatic vertebrate model organism in terms of the number of tested BPA alternatives as well as assessed endpoints and developmental stages (Supp. info. 6). Importantly, the collected studies show that BPA alternatives have not been systematically studied, i.e., studies that evaluated specific endpoints (e.g., effects related to endocrine disruption) across the models (fish vs. amphibians), or multiple chemicals, are largely lacking. The different chemicals have also been unequally investigated so far, with nearly 70 % of identified studies focusing on the top four BPA alternatives (BPS, BPF, BPAF, and TBBPA) (Fig. 5). Among the chemicals that have been examined, BPS has undergone the most comprehensive research in terms of both the quantity of studies conducted and the range of endpoints explored.

#### 4.5.3.1. Amphibian models.
Compared to fishes, much less studies on BPA and its alternatives are available in amphibians (Supp. info 6). The summary of their effects is visualized in Supp. info. 7. They are restricted to the order *Anura*, whereas no experiments using other amphibian orders, *Urodeles* and *Caecilians*, have been reported. The available studies focused mainly on investigations of BPA effects on sexual differentiation
<figure><img src="image_8.png" /><figcaption>**Fig. 5.** Depiction of the proportions of the collected research studies utilizing aquatic vertebrate models (fish and amphibians) that examined various BPA alternatives. The percentage number also reflects the studies when the specific bisphenol was studied in mixtures. The details of the studies are provided in Supp. info 6.</figcaption></figure>

and development during early larval phases (Oehlmann et al., 2009), demonstrating that already at 228 ng/L BPA can affect testicular development in *Xenopus laevis* tadpoles, by causing lacunae and testicular oocytes. In addition, BPA also affected growth and body weight; these were significantly increased in males exposed to concentrations of 2.28 µg/L BPA and above. This effect on growth could be explained by the induction of hepatic IGF-I (like growth factor 1) gene expression. Furthermore, a significant induction of VTG (vitellogenin) gene expression in the liver has been detected at 228 µg/L BPA.

Altogether, only five amphibian species, including *Xenopus laevis*, *Silurana tropicalis*, *Rhinella arenarum*, *Rana nigromaculatus*, and *Rana rugosa* have been used for investigating effects of BPA alternatives (BADGE, BPAF, BPB, BPF, BPS, DBD, TBBPA, TCBPA and TMBPA). The majority of these studies were using the established model organism *Xenopus laevis* to demonstrate endocrine disruptive effects. BADGE has been investigated only in *Rhinella arenarum* (Butler Wolkowicz et al., 2016) at stages ranging from stage 3 to 25 exposed continuously at concentrations from 0.0001 to 25 mg/L revealing LC₅₀ values for embryos and tadpoles with 0.04 and 2.2 mg/L, respectively. Pulsed exposures resulted in higher LC₅₀ values from 0.58 to 14.9 mg/L, depending on the exposed developmental stage. In addition, sub-lethal effects were also observed in exposed organisms. Continuous exposure induced cell dissociation during embryo development at concentrations equal to or above 0.5 mg/L. In larvae exposed to a concentration of 10 mg/L or higher, neurological alterations were observed; those exposed to 15 g/L showed narcotic effects. Furthermore, after a 168 h exposure to 5 mg/L, larvae exhibited signs of starvation, abnormal skin pigmentation, scare response to stimuli, and tail/axial flexures; while those exposed to 1 mg/L developed hydropsy and abnormal skin pigmentation. Regarding the observed sublethal effects after pulsed exposures, the main effects observed in embryos were microcephaly, hydropsy, axial flexures, and reduced body size, with a LOAEL of 1 mg/L.

In adult *R. nigromaculatus*, 14 d exposures to TCBPA and TBBPA ranging from 0.001 to 1 mg/L were performed and increased contents of alanine transaminase in the serum were observed in animals exposed from 0.001 to 0.1 mg/L of both compounds (Jia et al., 2022). ROS contents in liver as well as expression of cytochrome C increased significantly after exposure to TCBPA and TBBPA, at all concentrations tested, demonstrating metabolic effects. *R. rugosa* tadpoles were exposed to TBBPA for 9 days at 544; 54.4, and 5.44 µg/L in the presence or

absence of triiodothyronine (T3). T3-induced tail shrinking was inhibited by the lowest and highest tested concentrations (Kitamura et al., 2005), suggesting inhibitory or antagonistic modes of action concerning thyroid hormones. In addition, similar exposures with TBBPA, BPA, TCBPA, and TMBPA have been performed using *P. rugosa*, *P. tropicalis*, and *X. laevis* (Goto et al., 2006) (Kashiwagi et al., 2008). T3-induced tadpole tail shrinking in the wrinkled frog *R. rugosa* was suppressed by BPA, TBBPA, TCBPA, and TMBPA. BPA alternatives also inhibited spontaneous metamorphosis in the tropical clawed frog *S. tropicalis* controlled by endogenous circulating TH. These results indicate that BPA alternatives act as thyroid hormones (TH) antagonists. In *X. laevis* transgenic tadpoles carrying plasmid DNA containing TH response element (TRE) and 5'-upstream promoter region of the TH receptor (TR) βA1 gene linked to an enhanced green fluorescent protein (EGFP) gene, T3 induced a strong EGFP expression in the hind limbs. This induction was suppressed by BPA, TBBPA, TCBPA and TMBPA, suggesting BPA alternatives all act as antagonists to prevent the binding of T3 to TH receptor, resulting in inhibition of TR-mediated gene expression. In *X. laevis* tadpoles from stage NF 51, exposure for 21 d ranging from 2.5 to 500 µg/L TBBPA and short term exposure from 12 to 72 h with 0.1 and 1 µg/L T3 in the presence of 100, 250 and 500 µg/L TBBPA have been performed (Jagnytsch et al., 2006). TBBPA inhibited development and reduced hind limb length at 500 µg/L after 21 d. Short term exposure to 100, 250 and 500 µg/L TBBPA inhibited T3-stimulated increase of thyroidal biomarkers TRβ and b/ZIP gene expression, suggesting antagonistic modes of action on TH receptors as well.

Exposure to BPA and BPF ranging from 10⁻⁸ to 10⁻⁶ M in a *X. laevis* metamorphosis assay for 96 h (Niu et al., 2021) revealed T3 induced brain morphological remodeling coupled with cell proliferation and neuronal differentiation, whereas both BPA and BPF in general antagonized T3 induced changes in a dose dependent manner with weak or no effects of bisphenols exposure alone. BPA and BPF interfered with TH signaling in *X. laevis* brains, especially in the presence of thyroid hormones (TH), and subsequently affected TH-dependent brain development. Further, similar experiments (Zhu et al., 2018) investigating BPF with the T3 induced metamorphosis of *X. laevis* tadpoles exposed for 96 h at concentrations ranging from 10⁻⁸ to 10⁻⁵ M in the absence and presence of 10⁻⁹ M T3 revealed that higher concentrations of BPF (10⁻⁷ to 10⁻⁵ M) antagonized T3-induced TH response of gene transcription and morphological changes in a concentration-dependent manner whereas 10⁻⁹ M BPF exerted stimulatory effects on T-induced integral metamorphosis, displaying TH signaling disrupting effects with complicated concentration – response relationships. In the absence of T3, BPF inhibited development at metamorphic climax, but promoted pre- and pro-metamorphic development, displaying a developmental stage-dependent manner. Agonistic actions of BPF on Notch signaling in the intestine were observed, showing that BPF disrupts vertebrate development possibly via multiple pathways besides TH signaling.

*X. laevis* tadpoles at stages NF 45/46 (NF stages according to zahn et al. (2022) (Zahn et al., 2022) were exposed to BPB until stage NF 52 or until completion of metamorphosis from 10⁻⁸ to 10⁻⁶ M (to analyze whether potential effects of BPB on testis development are mediated by estrogen receptors). Therefore, tadpoles at stage NF 45/46 were exposed to 10⁻⁷ M BPB or the estrogen receptor antagonist ICI 182780, alone or in combination, until stage NF 52 (Li et al., 2022). Low concentrations of BPB disrupted testis differentiation partly via the estrogen receptor-mediated pathway and subsequently caused testicular dysgenesis after metamorphosis in *X. Laevis*. Estrogenic chemicals resulted in inhibition of testis differentiation, which might later lead to testicular dysgenesis in frogs indicating adverse impacts of BPB on amphibian reproduction

Early developmental effects of BPA and BPAF were investigated in *X. laevis* embryos at the two-cell stage 2 h post-fertilization (hpf) (Araúcio et al., 2019). BPA (10⁻⁶ to 5x10⁻⁵ M) and BPAF (3x10⁻⁹ to 2.5x10⁻⁵ M) caused disrupted cleavage divisions, slowed cytokinesis, and cellular dissociation within 1–6 h. Flexures of the spinal cord, shorter body axis/tail, craniofacial malformations, and significant
<figure><img src="image_8.png" /><figcaption>**Fig. 5.** Depiction of the proportions of the collected research studies utilizing aquatic vertebrate models (fish and amphibians) that examined various BPA alternatives. The percentage number also reflects the studies when the specific bisphenol was studied in mixtures. The details of the studies are provided in Supp. info 6.</figcaption></figure>

and development during early larval phases (Oehlmann et al., 2009), demonstrating that already at 228 ng/L BPA can affect testicular development in *Xenopus laevis* tadpoles, by causing lacunae and testicular oocytes. In addition, BPA also affected growth and body weight; these were significantly increased in males exposed to concentrations of 2.28 µg/L BPA and above. This effect on growth could be explained by the induction of hepatic IGF-I (like growth factor 1) gene expression. Furthermore, a significant induction of VTG (vitellogenin) gene expression in the liver has been detected at 228 µg/L BPA.

Altogether, only five amphibian species, including *Xenopus laevis*, *Silurana tropicalis*, *Rhinella arenarum*, *Rana nigromaculatus*, and *Rana rugosa* have been used for investigating effects of BPA alternatives (BADGE, BPAF, BPB, BPF, BPS, DBD, TBBPA, TCBPA and TMBPA). The majority of these studies were using the established model organism *Xenopus laevis* to demonstrate endocrine disruptive effects. BADGE has been investigated only in *Rhinella arenarum* (Butler Wolkowicz et al., 2016) at stages ranging from stage 3 to 25 exposed continuously at concentrations from 0.0001 to 25 mg/L revealing LC₅₀ values for embryos and tadpoles with 0.04 and 2.2 mg/L, respectively. Pulsed exposures resulted in higher LC₅₀ values from 0.58 to 14.9 mg/L, depending on the exposed developmental stage. In addition, sub-lethal effects were also observed in exposed organisms. Continuous exposure induced cell dissociation during embryo development at concentrations equal to or above 0.5 mg/L. In larvae exposed to a concentration of 10 mg/L or higher, neurological alterations were observed; those exposed to 15 g/L showed narcotic effects. Furthermore, after a 168 h exposure to 5 mg/L, larvae exhibited signs of starvation, abnormal skin pigmentation, scare response to stimuli, and tail/axial flexures; while those exposed to 1 mg/L developed hydropsy and abnormal skin pigmentation. Regarding the observed sublethal effects after pulsed exposures, the main effects observed in embryos were microcephaly, hydropsy, axial flexures, and reduced body size, with a LOAEL of 1 mg/L.

In adult *R. nigromaculatus*, 14 d exposures to TCBPA and TBBPA ranging from 0.001 to 1 mg/L were performed and increased contents of alanine transaminase in the serum were observed in animals exposed from 0.001 to 0.1 mg/L of both compounds (Jia et al., 2022). ROS contents in liver as well as expression of cytochrome C increased significantly after exposure to TCBPA and TBBPA, at all concentrations tested, demonstrating metabolic effects. *R. rugosa* tadpoles were exposed to TBBPA for 9 days at 544; 54.4, and 5.44 µg/L in the presence or

absence of triiodothyronine (T3). T3-induced tail shrinking was inhibited by the lowest and highest tested concentrations (Kitamura et al., 2005), suggesting inhibitory or antagonistic modes of action concerning thyroid hormones. In addition, similar exposures with TBBPA, BPA, TCBPA, and TMBPA have been performed using *P. rugosa*, *P. tropicalis*, and *X. laevis* (Goto et al., 2006) (Kashiwagi et al., 2008). T3-induced tadpole tail shrinking in the wrinkled frog *R. rugosa* was suppressed by BPA, TBBPA, TCBPA, and TMBPA. BPA alternatives also inhibited spontaneous metamorphosis in the tropical clawed frog *S. tropicalis* controlled by endogenous circulating TH. These results indicate that BPA alternatives act as thyroid hormones (TH) antagonists. In *X. laevis* transgenic tadpoles carrying plasmid DNA containing TH response element (TRE) and 5'-upstream promoter region of the TH receptor (TR) βA1 gene linked to an enhanced green fluorescent protein (EGFP) gene, T3 induced a strong EGFP expression in the hind limbs. This induction was suppressed by BPA, TBBPA, TCBPA and TMBPA, suggesting BPA alternatives all act as antagonists to prevent the binding of T3 to TH receptor, resulting in inhibition of TR-mediated gene expression. In *X. laevis* tadpoles from stage NF 51, exposure for 21 d ranging from 2.5 to 500 µg/L TBBPA and short term exposure from 12 to 72 h with 0.1 and 1 µg/L T3 in the presence of 100, 250 and 500 µg/L TBBPA have been performed (Jagnytsch et al., 2006). TBBPA inhibited development and reduced hind limb length at 500 µg/L after 21 d. Short term exposure to 100, 250 and 500 µg/L TBBPA inhibited T3-stimulated increase of thyroidal biomarkers TRβ and b/ZIP gene expression, suggesting antagonistic modes of action on TH receptors as well.

Exposure to BPA and BPF ranging from 10⁻⁸ to 10⁻⁶ M in a *X. laevis* metamorphosis assay for 96 h (Niu et al., 2021) revealed T3 induced brain morphological remodeling coupled with cell proliferation and neuronal differentiation, whereas both BPA and BPF in general antagonized T3 induced changes in a dose dependent manner with weak or no effects of bisphenols exposure alone. BPA and BPF interfered with TH signaling in *X. laevis* brains, especially in the presence of thyroid hormones (TH), and subsequently affected TH-dependent brain development. Further, similar experiments (Zhu et al., 2018) investigating BPF with the T3 induced metamorphosis of *X. laevis* tadpoles exposed for 96 h at concentrations ranging from 10⁻⁸ to 10⁻⁵ M in the absence and presence of 10⁻⁹ M T3 revealed that higher concentrations of BPF (10⁻⁷ to 10⁻⁵ M) antagonized T3-induced TH response of gene transcription and morphological changes in a concentration-dependent manner whereas 10⁻⁹ M BPF exerted stimulatory effects on T-induced integral metamorphosis, displaying TH signaling disrupting effects with complicated concentration – response relationships. In the absence of T3, BPF inhibited development at metamorphic climax, but promoted pre- and pro-metamorphic development, displaying a developmental stage-dependent manner. Agonistic actions of BPF on Notch signaling in the intestine were observed, showing that BPF disrupts vertebrate development possibly via multiple pathways besides TH signaling.

*X. laevis* tadpoles at stages NF 45/46 (NF stages according to zahn et al. (2022) (Zahn et al., 2022) were exposed to BPB until stage NF 52 or until completion of metamorphosis from 10⁻⁸ to 10⁻⁶ M (to analyze whether potential effects of BPB on testis development are mediated by estrogen receptors). Therefore, tadpoles at stage NF 45/46 were exposed to 10⁻⁷ M BPB or the estrogen receptor antagonist ICI 182780, alone or in combination, until stage NF 52 (Li et al., 2022). Low concentrations of BPB disrupted testis differentiation partly via the estrogen receptor-mediated pathway and subsequently caused testicular dysgenesis after metamorphosis in *X. Laevis*. Estrogenic chemicals resulted in inhibition of testis differentiation, which might later lead to testicular dysgenesis in frogs indicating adverse impacts of BPB on amphibian reproduction.

Early developmental effects of BPA and BPAF were investigated in *X. laevis* embryos at the two-cell stage 2 h post-fertilization (hpf) (Araúcio et al., 2019). BPA (10⁻⁶ to 5x10⁻⁵ M) and BPAF (3x10⁻⁹ to 2.5x10⁻⁵ M) caused disrupted cleavage divisions, slowed cytokinesis, and cellular dissociation within 1–6 h. Flexures of the spinal cord, shorter body axis/tail, craniofacial malformations, and significant
mortality occurred with doses of BPAF (LC₅₀ = 1.3x10⁻⁹ M). There were both shared and unique effects of all compounds, with BPAF having the greatest potency and toxicity (BPAF > BPA > estradiol). Further BPAF exposures ranging from 10⁻⁹ to 10⁻⁷ M have been performed using *X. laevis* tadpoles from stages NF 45/46 to 53 and 66 (Cai et al., 2020). All concentrations of BPAF caused changes in testicular morphology at

different developmental stages. Specifically, at stage NF 53, BPAF, like estradiol (E2), resulted in decreases in both, the size and the number of gonadal metameres (gonomeres), in testes, looking like ovaries. Some of the BPAF-treated testes remained segmented and even became discontinuous and fragmented at subsequent stages. Histological abnormalities were also observed in BPAF-treated testes, such as ovarian cavity at

<table><caption>**Fig. 6.** The overview of compiled research utilizing zebrafish as a model organism. The table shows what BPA alternatives are active with respect to the examined endpoints (Y- studied endpoint interacts with chemical, N- studied endpoint does NOT interact with chemical, Y/N- the studies show both trends). The studies used for this figure are described in detail in Supp. info. 6. The overview of compiled research for other studied fish models is presented in Supp. info. 7.</caption><thead><tr><th>Model organism</th><th>Impact on</th><th>Impacted endpoint</th><th>BPS</th><th>BPF</th><th>BPAF</th><th>BPB</th><th>BPAP</th><th>BPZ</th><th>BPC</th><th>DBP</th><th>BPE</th><th>BPM</th><th>BPP</th><th>BPFL</th><th>BADGE</th><th>TBPA</th><th>TCPA</th><th>TMBPA</th></tr></thead><tbody><tr><td rowspan="15">Zebrafish (Danio rerio)<br>Early life stages</td><td rowspan="15">Development</td><td>Hatch rate</td><td>Y/N</td><td>Y/N</td><td>Y/N</td><td>Y/N</td><td>Y/N</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Survival rate / mortality</td><td>Y/N</td><td>Y/N</td><td>Y/N</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Growth (length)</td><td>Y/N</td><td>Y/N</td><td>Y/N</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Yolk consumption</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Lipid composition</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Endoplasmatic reticulum</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Body weight</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Deformities / Morphology</td><td></td><td></td><td>Y</td><td>Y/N</td><td>Y/N</td><td>N</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Loss of pigmentation</td><td></td><td></td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td rowspan="2">Oxidative stress</td><td>ROS</td><td>Y</td><td>Y</td><td>Y/N</td><td>Y</td><td>N</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Enzymes activity</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td rowspan="2">Immunity</td><td>Immunity markers</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Locomotion activity</td><td>Y</td><td>Y</td><td>Y</td><td>Y/N</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td rowspan="10">Neurotoxicity</td><td>Neurogenesis/Neurodevelopment</td><td>Y/N</td><td>Y/N</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Neurotransmitters</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Brain structure</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Behavior</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Ache activity</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Microglia, Astrocytes</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Neuroinflammatory response</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td rowspan="10">Endocrine disruption</td><td>Thyroid hormones</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Estrogen Receptor</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Insulin signaling</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Testosterone</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Estrogen (ER alfa)</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Vitellogenin (Vtg)</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Antiestrogen effect</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Sensory endpoints</td><td>Retinal structure</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>GI tract</td><td>Intestinal Integrity</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Apoptosis</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td rowspan="2">Cardiac endpoints</td><td>Heart rate, Heartbeats</td><td>Y</td><td>Y</td><td>Y/N</td><td>Y</td><td>N</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Loss of blood vessels</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Metabolic</td><td>Glucometabolic</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td rowspan="15">Zebrafish (Danio rerio)<br>Adult stages</td><td rowspan="15">Microbiome</td><td>Community structure</td><td>Y</td><td>Y</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td></tr><tr><td>Diversity indexes</td><td>Y</td><td>Y/N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td></tr><tr><td>Microbiome function</td><td>Y</td><td>Y</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td></tr><tr><td rowspan="2">Development</td><td>Growth (length)</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Body weight</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td rowspan="3">Neurotoxicity</td><td>Locomotion activity</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Social behavior</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Anxiety response</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td rowspan="7">Endocrine disruption</td><td>Estrogen receptor levels</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Steroidogenic enzymes</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Vitellogenin (Vtg)</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>17b estradiol (E2)</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Thyroxine</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Triiodothyronine</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Testosterone</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td rowspan="2">Oxidative stress</td><td>ROS / oxidative stress</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Enzymes activity</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td rowspan="5">DNA damage</td><td>DNA methylation</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Reproduction</td><td>Reproductive capacity (males)</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Gonadosomatic index</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Sperm count</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Egg production</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td rowspan="3">Transgenerational effect</td><td>Ovarian and testicular histology</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Alter social behavior</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Hepatotoxicity</td><td>Hepatosomatic index</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td></td><td>Fatty acid synthesis/degradation</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td></td><td>Fat / TAG accumulation</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr></tbody></table>
mortality occurred with doses of BPAF (LC₅₀ = 1.3x10⁻⁹ M). There were both shared and unique effects of all compounds, with BPAF having the greatest potency and toxicity (BPAF > BPA > estradiol). Further BPAF exposures ranging from 10⁻⁹ to 10⁻⁷ M have been performed using X. laevis tadpoles from stages NF 45/46 to 53 and 66 (Cai et al., 2020). All concentrations of BPAF caused changes in testicular morphology at

different developmental stages. Specifically, at stage NF 53, BPAF, like estradiol (E2), resulted in decreases in both, the size and the number of gonadal metameres (gonomeres), in testes, looking like ovaries. Some of the BPAF-treated testes remained segmented and even became discontinuous and fragmented at subsequent stages. Histological abnormalities were also observed in BPAF-treated testes, such as ovarian cavity at

<table><caption>**Fig. 6.** The overview of compiled research utilizing zebrafish as a model organism. The table shows what BPA alternatives are active with respect to the examined endpoints (Y- studied endpoint interacts with chemical, N- studied endpoint does NOT interact with chemical, Y/N- the studies show both trends). The studies used for this figure are described in detail in Supp. info. 6. The overview of compiled research for other studied fish models is presented in Supp. info. 7.</caption><thead><tr><th>Model organism</th><th>Impact on</th><th>Impacted endpoint</th><th>BPS</th><th>BPF</th><th>BPAF</th><th>BPB</th><th>BPAP</th><th>BPZ</th><th>BPC</th><th>DBP</th><th>BPE</th><th>BPM</th><th>BPP</th><th>BPFL</th><th>BADGE</th><th>TBPA</th><th>TCPA</th><th>TMBPA</th></tr></thead><tbody><tr><td rowspan="12">Zebrafish (Danio rerio)<br>Early life stages</td><td rowspan="12">Development</td><td>Hatch rate</td><td>Y/N</td><td>Y/N</td><td>Y/N</td><td>Y/N</td><td>Y/N</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Survival rate / mortality</td><td>Y/N</td><td>Y/N</td><td>Y/N</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Growth (length)</td><td>Y/N</td><td>Y/N</td><td>Y/N</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Yolk consumption</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Lipid composition</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Endoplasmatic reticulum</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Body weight</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Deformities / Morphology</td><td></td><td></td><td>Y</td><td>Y/N</td><td>Y/N</td><td>N</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Loss of pigmentation</td><td></td><td></td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td rowspan="2">Oxidative stress</td><td>ROS</td><td>Y</td><td>Y</td><td>Y/N</td><td>Y</td><td>N</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Enzymes activity</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td rowspan="2">Immunity</td><td>Immunity markers</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Locomotion activity</td><td>Y</td><td>Y</td><td>Y</td><td>Y/N</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td rowspan="12">Neurotoxicity</td><td>Neurogenesis/Neurodevelopment</td><td>Y/N</td><td>Y/N</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Neurotransmitters</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Brain structure</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Behavior</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Ache activity</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Microglia, Astrocytes</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Neuroinflammatory response</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td rowspan="12">Endocrine disruption</td><td>Thyroid hormones</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Estrogen Receptor</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Insulin signaling</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Testosterone</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Estrogen (ER alfa)</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Vitellogenin (Vtg)</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Antiestrogen effect</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Sensory endpoints</td><td>Retinal structure</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>GI tract</td><td>Intestinal Integrity</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Apoptosis</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td rowspan="2">Cardiac endpoints</td><td>Heart rate, Heartbeats</td><td>Y</td><td>Y</td><td>Y/N</td><td>Y</td><td>N</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Loss of blood vessels</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td rowspan="12">Zebrafish (Danio rerio)<br>Adult stages</td><td rowspan="12">Microbiome</td><td>Glucometabolic</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Community structure</td><td>Y</td><td>Y</td><td>Y</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td></tr><tr><td>Diversity indexes</td><td>Y</td><td>Y/N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td></tr><tr><td>Microbiome function</td><td>Y</td><td>Y</td><td>Y</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td><td>N</td></tr><tr><td rowspan="2">Development</td><td>Growth (length)</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Body weight</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td rowspan="2">Neurotoxicity</td><td>Locomotion activity</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Social behavior</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td rowspan="12">Endocrine disruption</td><td>Estrogen receptor levels</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Steroidogenic enzymes</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Vitellogenin (vtg)</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>17b estradiol (E2)</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Thyroxine</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Triiodothyronine</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Testosterone</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>ROS / oxidative stress</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Enzymes activity</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td rowspan="2">DNA damage</td><td>DNA methylation</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Reproduction</td><td>Reproductive capacity (males)</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td rowspan="12">Transgenerational effect</td><td rowspan="12">Hepatotoxicity</td><td>Gonadosomatic index</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Sperm count</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Egg production</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Ovarian and testicular histology</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>DNA methylation in offsprings</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Alter social behavior</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Hepatosomatic index</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Fatty acid synthesis/degradation</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr><tr><td>Fat / TAG accumulation</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td><td>Y</td></tr></tbody></table>
stages NF 53 and 66 and poorly developed seminiferous tubules 8 weeks post metamorphosis. At the molecular level, BPAF inhibited expression of genes highly expressed in testes of males at stage NF 53. Correspondingly, BPAF, like E2, inhibited cell proliferation in testes at stage NF 50. All results show that low concentrations of BPAF inhibited testicular differentiation and subsequent development in *X. laevis* by feminizing effects.

*X. laevis* tadpoles at stage NF 52 were exposed to BPA or BPF ranging from $10^{-8}$ to $10^{-6}$ M for 5 days (Zhu et al., 2018). BPA and BPF both significantly elevated notch-related gene expression in a dose dependent manner. Cell proliferation and intestinal histological changes were observed. The study demonstrated that both, BPA and BPF, activate Notch signaling and subsequently disrupt intestinal development

In summary, the rare studies dealing with BPA alternatives in amphibians suggest that BPA related compounds might similarly affect sexual differentiation by estrogenic or antiandrogenic modes of action. In addition, some compounds might also impact TH receptors in an antagonistic manner or even affect multiple yet unraveled metabolic pathways. Despite the fact that investigated BPA alternatives are often less effective than BPA or cause adverse effects, especially at very early developmental stages, only at relatively high concentrations, it has to be concluded that they are endocrine disruptive chemicals (EDCs) being effective in disrupting reproduction, development and metabolism.

#### 4.5.3.2. Fish models.
Upon reviewing the existing studies, we found that twelve different fish models were utilized to assess and characterize the bioactivity of BPA alternatives. In majority, the studies focused on examining the effects on early stages of zebrafish (Supp. info. 6). This preference for using zebrafish as a model aligns with other toxicological and pharmaceutical research, where the zebrafish model has a dominant presence due to its favorable experimental features as well as conservation of a significant number of morphological and physiological processes in vertebrate organogenesis (MacRae and Peterson, 2015). With respect to BPA alternative research, the other fish species were typically employed to study only one or two chemicals. The studied BPA alternatives and specific endpoints evaluated in zebrafish are summarized in Supp. info 7.

Numerous studies have shown that BPA functions as an endocrine disruptor, capable of perturbing hormone levels and metabolism in vertebrates (Pelch et al., 2019). As a result, the primary emphasis of research on BPA alternatives has been on assessing their potential to interfere with the endocrine system. Many of the BPA alternatives demonstrate their influence on fish by affecting gene and hormone levels involved in reproduction, development, and growth. These studies reveal significant effects on gene expression associated with steroidogenesis and sex hormone production. Induction of vitellogenin, a precursor to egg yolk typically synthesized in females, was also observed in males, confirming the estrogenic effects of specific BPA alternatives (Fig. 6). In addition to vitellogenin expression and production, *in vitro* and transgenic fish models have been also utilized to investigate the interaction with estrogen receptors and the modulation of estrogen-specific pathways. It has been shown that BPA and some of its substitutes (BPF, BPS, BPC, BPAF) have the capacity to interact with the different zebrafish ER subtypes in a selective manner. Hence in human or zebrafish cells expressing the zebrafish ER-subtypes alpha, beta1 or Beta 2, BPA and some bisphenol substitutes efficiently transactivate zebrafish estrogen subtypes, with a differential selectivity depending on the bisphenol tested and the zfER subtypes (Le Fol et al., 2017; Pinto et al., 2019). At the *in vivo* level the use of estrogeno-sensitive transgenic zebrafish embryo models such as tg(5xERE:GFP) and/or tg(cyp19a1b: GFP) allowed to demonstrate the capacity of BPA, BPS, BPF,BPC and BPAF to induce ER-signaling pathway in different tissue expressing ERs such as the heart, the liver and the brain (Le Fol et al., 2017; Pinto et al., 2019; Moreman et al., 2017). The ability of bisphenols to induce brain aromatase in zebrafish seems a common feature as all the bisphenols

tested up now are active in tg (cyp19a1b:GFP) (Christophe et al., under revision) (Christophe, xxxx). Disruption of brain aromatase by bisphenols could play a role in their behavioral effect but a recent study showed that the behavioral changes (e.g. locomotor effect) measured in bisphenol-exposed zebrafish could not be linked to their potency to modulate brain aromatase expression (Blanc-Legendre et al., 2023). These studies highlighted the relevance of combining cellular and whole-organism bioassays in zebrafish for the hazard assessment of bisphenols by providing relevant data on their mode of action on endocrine pathway and behavioral effect. From a regulatory point of view, the data obtained using these assays can not be used for risk assessment of these bisphenols. However, the reported mode of action and effect being similar to BPA they should be considered to accelerate the process of identifying these substitutes as endocrine disruptors.

In some cases (e.g. BPS), we observed that there might be no need to evaluate its endocrine disruptive effects in any further detail, as the extensive evidence already available shows endocrine disruption potency on multiple levels. Thus, our findings highlight not only the importance of further toxicological research for less studied BPA alternatives but also demonstrate that some BPA alternatives have already been studied in sufficient detail to allow a reliable determination of their associated hazards, specifically connected with endocrine disruption.

BPA exposure has been shown to decrease the activity of antioxidant enzymes, such as superoxide dismutase (SOD) and catalase, which are responsible for neutralizing ROS (reactive oxygen species) and maintaining cellular redox balance (Pradhan et al., 2021). As a consequence, the accumulation of ROS overwhelms the antioxidant defenses, resulting in oxidative stress and the potential disruption of various cellular processes in zebrafish. Oxidative stress appears to be a conserved effect observed in response not only to BPA but also to BPA alternatives. Similar to BPA, these alternatives can modulate levels of antioxidant enzymes and induce oxidative stress in various organisms, including zebrafish (Fig. 6.) and other investigated fish models (Supp. info 7) (e.g. African catfish, rainbow trout or marine medaka). Thus, oxidative stress appears to be a common pathway through which both BPA and its alternatives exert their toxic effects. This emphasizes the importance of considering oxidative stress as a critical aspect when evaluating the potential risks associated with these compounds.

With respect to BPA alternatives, the various adverse outcomes have been studied in a rather non-systematic way. For many of the chemicals, and in contrast to endocrine disruptive potency, many of the endpoints are studied less frequently, for example, the immunity and microbiome. The microbiome and immune system, although crucial aspects of an organism's health, have been overlooked as endpoints in the study of BPA alternatives. While the effects of BPA on the microbiome and immunity have been investigated (Xu et al., 2013; McDonough et al., 2021), only one study examined the impact of BPA alternatives on these endpoints (Catron et al., 2019). The microbiome, consisting of the diverse community of microorganisms residing in the body, plays a vital role in various physiological processes, including immune function (Sehnal et al., 2021). Disruptions to the microbiome can have far-reaching consequences for overall health and immune system regulation (Adamovsky et al., 2018). Thus, by neglecting to assess the effects of BPA alternatives on the microbiome and immune system, our understanding of their potential health risks on aquatic vertebrates remains incomplete. It is essential to broaden the scope of research to include these critical endpoints to comprehensively evaluate the environmental safety and biological impacts of BPA alternatives.

In conclusion, the summary of the effects of BPA alternatives on aquatic vertebrates indicates that some of the chemicals can be hazardous, sharing the mode of action with BPA (e.g. deregulation of endocrine system and induction of oxidative stress). Future research should focus on exploring overlooked endpoints, as well as investigating the long-term effects and potential ecological implications of less studied BPA alternatives on fish populations and aquatic ecosystems. Further, for certain BPA alternatives (e.g. BPS), the existing scientific
stages NF 53 and 66 and poorly developed seminiferous tubules 8 weeks post metamorphosis. At the molecular level, BPAF inhibited expression of genes highly expressed in testes of males at stage NF 53. Correspondingly, BPAF, like E2, inhibited cell proliferation in testes at stage NF 50. All results show that low concentrations of BPAF inhibited testicular differentiation and subsequent development in *X. laevis* by feminizing effects.

*X. laevis* tadpoles at stage NF 52 were exposed to BPA or BPF ranging from $10^{-8}$ to $10^{-6}$ M for 5 days (Zhu et al., 2018). BPA and BPF both significantly elevated notch-related gene expression in a dose dependent manner. Cell proliferation and intestinal histological changes were observed. The study demonstrated that both, BPA and BPF, activate Notch signaling and subsequently disrupt intestinal development

In summary, the rare studies dealing with BPA alternatives in amphibians suggest that BPA related compounds might similarly affect sexual differentiation by estrogenic or antiandrogenic modes of action. In addition, some compounds might also impact TH receptors in an antagonistic manner or even affect multiple yet unraveled metabolic pathways. Despite the fact that investigated BPA alternatives are often less effective than BPA or cause adverse effects, especially at very early developmental stages, only at relatively high concentrations, it has to be concluded that they are endocrine disruptive chemicals (EDCs) being effective in disrupting reproduction, development and metabolism.

#### 4.5.3.2. Fish models.
Upon reviewing the existing studies, we found that twelve different fish models were utilized to assess and characterize the bioactivity of BPA alternatives. In majority, the studies focused on examining the effects on early stages of zebrafish (Supp. info. 6). This preference for using zebrafish as a model aligns with other toxicological and pharmaceutical research, where the zebrafish model has a dominant presence due to its favorable experimental features as well as conservation of a significant number of morphological and physiological processes in vertebrate organogenesis (MacRae and Peterson, 2015). With respect to BPA alternative research, the other fish species were typically employed to study only one or two chemicals. The studied BPA alternatives and specific endpoints evaluated in zebrafish are summarized in Supp. info 7.

Numerous studies have shown that BPA functions as an endocrine disruptor, capable of perturbing hormone levels and metabolism in vertebrates (Pelch et al., 2019). As a result, the primary emphasis of research on BPA alternatives has been on assessing their potential to interfere with the endocrine system. Many of the BPA alternatives demonstrate their influence on fish by affecting gene and hormone levels involved in reproduction, development, and growth. These studies reveal significant effects on gene expression associated with steroidogenesis and sex hormone production. Induction of vitellogenin, a precursor to egg yolk typically synthesized in females, was also observed in males, confirming the estrogenic effects of specific BPA alternatives (Fig. 6). In addition to vitellogenin expression and production, *in vitro* and transgenic fish models have been also utilized to investigate the interaction with estrogen receptors and the modulation of estrogen-specific pathways. It has been shown that BPA and some of its substitutes (BPF, BPS, BPC, BPAF) have the capacity to interact with the different zebrafish ER subtypes in a selective manner. Hence in human or zebrafish cells expressing the zebrafish ER-subtypes alpha, beta1 or Beta 2, BPA and some bisphenol substitutes efficiently transactivate zebrafish estrogen subtypes, with a differential selectivity depending on the bisphenol tested and the zfER subtypes (Le Fol et al., 2017; Pinto et al., 2019). At the *in vivo* level the use of estrogeno-sensitive transgenic zebrafish embryo models such as tg(5xERE:GFP) and/or tg(cyp19a1b: GFP) allowed to demonstrate the capacity of BPA, BPS, BPF, BPC and BPAF to induce ER-signaling pathway in different tissue expressing ERs such as the heart, the liver and the brain (Le Fol et al., 2017; Pinto et al., 2019; Moreman et al., 2017). The ability of bisphenols to induce brain aromatase in zebrafish seems a common feature as all the bisphenols

tested up now are active in tg (cyp19a1b:GFP) (Christophe et al., under revision) (Christophe, xxxx). Disruption of brain aromatase by bisphenols could play a role in their behavioral effect but a recent study showed that the behavioral changes (e.g. locomotor effect) measured in bisphenol-exposed zebrafish could not be linked to their potency to modulate brain aromatase expression (Blanc-Legendre et al., 2023). These studies highlighted the relevance of combining cellular and whole-organism bioassays in zebrafish for the hazard assessment of bisphenols by providing relevant data on their mode of action on endocrine pathway and behavioral effect. From a regulatory point of view, the data obtained using these assays can not be used for risk assessment of these bisphenols. However, the reported mode of action and effect being similar to BPA they should be considered to accelerate the process of identifying these substitutes as endocrine disruptors.

In some cases (e.g. BPS), we observed that there might be no need to evaluate its endocrine disruptive effects in any further detail, as the extensive evidence already available shows endocrine disruption potency on multiple levels. Thus, our findings highlight not only the importance of further toxicological research for less studied BPA alternatives but also demonstrate that some BPA alternatives have already been studied in sufficient detail to allow a reliable determination of their associated hazards, specifically connected with endocrine disruption.

BPA exposure has been shown to decrease the activity of antioxidant enzymes, such as superoxide dismutase (SOD) and catalase, which are responsible for neutralizing ROS (reactive oxygen species) and maintaining cellular redox balance (Pradhan et al., 2021). As a consequence, the accumulation of ROS overwhelms the antioxidant defenses, resulting in oxidative stress and the potential disruption of various cellular processes in zebrafish. Oxidative stress appears to be a conserved effect observed in response not only to BPA but also to BPA alternatives. Similar to BPA, these alternatives can modulate levels of antioxidant enzymes and induce oxidative stress in various organisms, including zebrafish (Fig. 6) and other investigated fish models (Supp. info 7) (e.g. African catfish, rainbow trout or marine medaka). Thus, oxidative stress appears to be a common pathway through which both BPA and its alternatives exert their toxic effects. This emphasizes the importance of considering oxidative stress as a critical aspect when evaluating the potential risks associated with these compounds.

With respect to BPA alternatives, the various adverse outcomes have been studied in a rather non-systematic way. For many of the chemicals, and in contrast to endocrine disruptive potency, many of the endpoints are studied less frequently, for example, the immunity and microbiome. The microbiome and immune system, although crucial aspects of an organism's health, have been overlooked as endpoints in the study of BPA alternatives. While the effects of BPA on the microbiome and immunity have been investigated (Xu et al., 2013; McDonough et al., 2021), only one study examined the impact of BPA alternatives on these endpoints (Catron et al., 2019). The microbiome, consisting of the diverse community of microorganisms residing in the body, plays a vital role in various physiological processes, including immune function (Sehnal et al., 2021). Disruptions to the microbiome can have far-reaching consequences for overall health and immune system regulation (Adamovsky et al., 2018). Thus, by neglecting to assess the effects of BPA alternatives on the microbiome and immune system, our understanding of their potential health risks on aquatic vertebrates remains incomplete. It is essential to broaden the scope of research to include these critical endpoints to comprehensively evaluate the environmental safety and biological impacts of BPA alternatives.

In conclusion, the summary of the effects of BPA alternatives on aquatic vertebrates indicates that some of the chemicals can be hazardous, sharing the mode of action with BPA (e.g. deregulation of endocrine system and induction of oxidative stress). Future research should focus on exploring overlooked endpoints, as well as investigating the long-term effects and potential ecological implications of less studied BPA alternatives on fish populations and aquatic ecosystems. Further, for certain BPA alternatives (e.g. BPS), the existing scientific
studies have already provided extensive evidence regarding their potential dangers to the environment. Therefore, there is a strong rationale to assert that conducting additional toxicological studies on these alternatives would be redundant and unnecessary. Decision makers are now faced with a responsibility to take swift and decisive action based on the available research data. It is imperative that they prioritize the protection of the environment and public health by implementing appropriate measures to regulate or restrict the use of identified hazardous BPA alternatives. By taking proactive steps based on the existing knowledge, decision makers can effectively mitigate the potential risks associated with these substances and safeguard the well-being of ecosystems and living organisms.

## 5. Conclusion and future perspectives

BPA and BPA alternatives are chemicals commonly found in all environmental compartments. Although BPA is currently regulated and its use is restricted in many products, BPA can still be found in the environment at a relatively high level in comparison with its alternatives. However, it can be expected that levels of BPA alternatives will dominate in future. Therefore, the monitoring and surveillance of water and sediment quality should be enhanced to include measurements of BPA alternatives to monitor their levels and ensure development of timely mitigation strategies.

The review on QSAR brings attention to two primary concerns. Firstly, there is a notable inadequacy in the representation of bisphenols within the training datasets, coupled with a scarcity of experimental data available for compounds related to bisphenols. Secondly, there exists a considerable demand for comprehensive *in silico* methods that encompass the entire spectrum of bisphenols, including their derivatives and alternative compounds. This demand arises from the significance of certain endpoints, which also carry regulatory importance. Within the toxicokinetic aspect of this review, the dynamic models analyzed suggest potential variations in ADME (absorption, distribution, metabolism, and excretion) processes among bisphenol analogues. This calls for further research to validate these findings. On the toxicodynamic aspect, this review underlines the necessity for predictive models capable of assessing the effects of bisphenols across various levels of biological organization, ranging from individual organisms to entire ecosystems.

When comparing BPA to its alternatives, it becomes evident that there has been less extensive research into the assessment of the biological activities of BPA alternatives. However, these BPA alternatives are indeed present in the environment, their effects have not been adequately examined through controlled experiments that prioritize specific bisphenols, such as PBS. The available toxicological studies nevertheless cover experimental models from many parts of the phylogenetic tree, starting with microorganisms to vertebrates. In terms of the impact on microbiome, the studies emphasize the substantial impact that both BPA and its alternatives can exert on microbial organisms and communities, disrupting their physiology, operations, and composition. This could potentially have an impact for hosts and the functions of ecosystems and the services they provide. The chapters promote the idea of including microbes and their communities as integral components of conventional hazard assessment processes, specifically for BPA alternatives. Our investigation also delved into the significance of exploring the endocrine-disrupting effects of both BPA and its alternatives on invertebrates, with a particular focus on arthropods and insects. We stressed the importance of conducting more comprehensive assessments that go beyond the traditional reproductive studies. These investigations reveal that BPA and its alternatives can indeed interfere with the endocrine systems of insects, particularly by disrupting the ecdysis process and potentially causing genotoxic effects. This underscores the necessity for utilizing advanced molecular, analytical, and omics techniques to pinpoint molecular markers that can be used to evaluate potential adverse impacts on the insect endocrine systems. Furthermore, it highlights the need for developing new AOPs that are specifically related

to the ecdysis processes in invertebrates.

The presence of BPA and its alternatives in mollusks found in the environment stresses the need to incorporate mollusks into future research efforts. Although the effects of BPA on mollusks have been thoroughly examined, there is a shortage of research regarding the impact of BPA alternatives on these organisms. Consequently, it is vital to conduct additional investigations in this domain, particularly focusing on the early life stages of mollusks. Ascidians were another class of organisms used to study the endocrine effect of BPA. However, none BPA alternatives were studied (only BPA), using this model, the studies show ascidians as valuable model organisms for studying endocrine-disrupting chemicals due to their close evolutionary proximity to vertebrates and various advantageous characteristics for laboratory studies. Similarly, crustaceans (e.g., Daphnia spp., Artemia spp.) are a valuable model organism for evaluating the impacts of environmental pollutants, including endocrine-disrupting chemicals. It might be concluded that most of the available data refers to BPA and the information about the toxicity of BPA alternatives is very inconsistent and limited to specific endpoints including neuro-endocrine axis. Earthworms (phylum Nematoda) provide valuable insights into the toxic effects of BPA and its alternatives. BPA alternatives impact a broad spectrum of biological systems in earthworms exhibiting effects in growth, reproduction, behavior, and metabolism. Interestingly, *C. elegans* were successfully used for investigation of obesogenic potency of BPA alternatives such as BPS. Cnidaria (Hydra spp.) display varying degrees of sensitivity to bisphenols like BPA and BADGE which shows that BPA alternatives can be sensed by this family of organisms. This underscores the need for further research on the ecotoxicity of bisphenols in these organisms, considering their ecological significance in freshwater ecosystems.

The impact of BPA alternatives on higher organisms were also studied using vertebrate models. For studies employing vertebrate models, we found that current studies cover only selected BPA alternatives and thus have not been systematically studied (nearly 70 % of identified studies focused on the top four BPA alternatives (BPS, BPF, BPAF, and TBBPA)). Specifically, the existing studies on amphibians, particularly in the order Anura, demonstrate that bisphenols, including BPA and its alternatives, can disrupt sexual differentiation, development, metabolism and endocrine systems in various amphibian species. Similarly for fish studies, research primarily focused on impact on endocrine systems. In fish, BPA alternatives can disrupt gene and hormone levels associated with reproduction, development, and growth, and induce oxidative stress.

In summary, to ensure timely mitigation strategies, future monitoring and surveillance of the quality of the environment should be enhanced to include measurements of BPA alternatives and monitor their levels closely. The biological activity has been studied for BPA alternatives, but in a non-systematic way and prioritizing only a limited number of chemicals. Nonetheless, for several common BPA alternatives, extensive scientific research has already provided substantial evidence regarding their potential harm to the environment. The collective assessment of BPA alternatives suggests that certain chemicals may pose hazards, mirroring the mode of action seen with BPA, such as endocrine system disruption and oxidative stress induction. Therefore, it is reasonable to argue that carrying out further toxicological investigations on specific well-studied BPA alternatives would be superfluous and unwarranted. In terms of studied endpoints, several key biological systems are overlooked including the impact on immunity. Future research should emphasize unexplored aspects, the possible adverse effects caused by the exposure to mixtures of bisphenols at realistic environmental concentrations and the long-term consequences on ecosystem.

In the frame of project PARC (work package 5 (WP5)) and more specifically in the frame of internal project "BPA alternatives and associated mixtures (Data gaps and NAM development)" we will focus on the investigation of potential adverse effects of certain individual BPA alternatives and "real-life" mixtures on different species to fill the
studies have already provided extensive evidence regarding their potential dangers to the environment. Therefore, there is a strong rationale to assert that conducting additional toxicological studies on these alternatives would be redundant and unnecessary. Decision makers are now faced with a responsibility to take swift and decisive action based on the available research data. It is imperative that they prioritize the protection of the environment and public health by implementing appropriate measures to regulate or restrict the use of identified hazardous BPA alternatives. By taking proactive steps based on the existing knowledge, decision makers can effectively mitigate the potential risks associated with these substances and safeguard the well-being of ecosystems and living organisms.

## 5. Conclusion and future perspectives

BPA and BPA alternatives are chemicals commonly found in all environmental compartments. Although BPA is currently regulated and its use is restricted in many products, BPA can still be found in the environment at a relatively high level in comparison with its alternatives. However, it can be expected that levels of BPA alternatives will dominate in future. Therefore, the monitoring and surveillance of water and sediment quality should be enhanced to include measurements of BPA alternatives to monitor their levels and ensure development of timely mitigation strategies.

The review on QSAR brings attention to two primary concerns. Firstly, there is a notable inadequacy in the representation of bisphenols within the training datasets, coupled with a scarcity of experimental data available for compounds related to bisphenols. Secondly, there exists a considerable demand for comprehensive *in silico* methods that encompass the entire spectrum of bisphenols, including their derivatives and alternative compounds. This demand arises from the significance of certain endpoints, which also carry regulatory importance. Within the toxicokinetic aspect of this review, the dynamic models analyzed suggest potential variations in ADME (absorption, distribution, metabolism, and excretion) processes among bisphenol analogues. This calls for further research to validate these findings. On the toxicodynamic aspect, this review underlines the necessity for predictive models capable of assessing the effects of bisphenols across various levels of biological organization, ranging from individual organisms to entire ecosystems.

When comparing BPA to its alternatives, it becomes evident that there has been less extensive research into the assessment of the biological activities of BPA alternatives. However, these BPA alternatives are indeed present in the environment, their effects have not been adequately examined through controlled experiments that prioritize specific bisphenols, such as PBS. The available toxicological studies nevertheless cover experimental models from many parts of the phylogenetic tree, starting with microorganisms to vertebrates. In terms of the impact on microbiome, the studies emphasize the substantial impact that both BPA and its alternatives can exert on microbial organisms and communities, disrupting their physiology, operations, and composition. This could potentially have an impact for hosts and the functions of ecosystems and the services they provide. The chapters promote the idea of including microbes and their communities as integral components of conventional hazard assessment processes, specifically for BPA alternatives. Our investigation also delved into the significance of exploring the endocrine-disrupting effects of both BPA and its alternatives on invertebrates, with a particular focus on arthropods and insects. We stressed the importance of conducting more comprehensive assessments that go beyond the traditional reproductive studies. These investigations reveal that BPA and its alternatives can indeed interfere with the endocrine systems of insects, particularly by disrupting the ecdysis process and potentially causing genotoxic effects. This underscores the necessity for utilizing advanced molecular, analytical, and omics techniques to pinpoint molecular markers that can be used to evaluate potential adverse impacts on the insect endocrine systems. Furthermore, it highlights the need for developing new AOPs that are specifically related

to the ecdysis processes in invertebrates.

The presence of BPA and its alternatives in mollusks found in the environment stresses the need to incorporate mollusks into future research efforts. Although the effects of BPA on mollusks have been thoroughly examined, there is a shortage of research regarding the impact of BPA alternatives on these organisms. Consequently, it is vital to conduct additional investigations in this domain, particularly focusing on the early life stages of mollusks. Ascidians were another class of organisms used to study the endocrine effect of BPA. However, none BPA alternatives were studied (only BPA), using this model, the studies show ascidians as valuable model organisms for studying endocrine-disrupting chemicals due to their close evolutionary proximity to vertebrates and various advantageous characteristics for laboratory studies. Similarly, crustaceans (e.g., Daphnia spp., Artemia spp.) are a valuable model organism for evaluating the impacts of environmental pollutants, including endocrine-disrupting chemicals. It might be concluded that most of the available data refers to BPA and the information about the toxicity of BPA alternatives is very inconsistent and limited to specific endpoints including neuro-endocrine axis. Earthworms (phylum Nematoda) provide valuable insights into the toxic effects of BPA and its alternatives. BPA alternatives impact a broad spectrum of biological systems in earthworms exhibiting effects in growth, reproduction, behavior, and metabolism. Interestingly, *C. elegans* were successfully used for investigation of obesogenic potency of BPA alternatives such as BPS. Cnidaria (Hydra spp.) display varying degrees of sensitivity to bisphenols like BPA and BADGE which shows that BPA alternatives can be sensed by this family of organisms. This underscores the need for further research on the ecotoxicity of bisphenols in these organisms, considering their ecological significance in freshwater ecosystems.

The impact of BPA alternatives on higher organisms were also studied using vertebrate models. For studies employing vertebrate models, we found that current studies cover only selected BPA alternatives and thus have not been systematically studied (nearly 70 % of identified studies focused on the top four BPA alternatives (BPS, BPF, BPAF, and TBBPA)). Specifically, the existing studies on amphibians, particularly in the order Anura, demonstrate that bisphenols, including BPA and its alternatives, can disrupt sexual differentiation, development, metabolism and endocrine systems in various amphibian species. Similarly for fish studies, research primarily focused on impact on endocrine systems. In fish, BPA alternatives can disrupt gene and hormone levels associated with reproduction, development, and growth, and induce oxidative stress.

In summary, to ensure timely mitigation strategies, future monitoring and surveillance of the quality of the environment should be enhanced to include measurements of BPA alternatives and monitor their levels closely. The biological activity has been studied for BPA alternatives, but in a non-systematic way and prioritizing only a limited number of chemicals. Nonetheless, for several common BPA alternatives, extensive scientific research has already provided substantial evidence regarding their potential harm to the environment. The collective assessment of BPA alternatives suggests that certain chemicals may pose hazards, mirroring the mode of action seen with BPA, such as endocrine system disruption and oxidative stress induction. Therefore, it is reasonable to argue that carrying out further toxicological investigations on specific well-studied BPA alternatives would be superfluous and unwarranted. In terms of studied endpoints, several key biological systems are overlooked including the impact on immunity. Future research should emphasize unexplored aspects, the possible adverse effects caused by the exposure to mixtures of bisphenols at realistic environmental concentrations and the long-term consequences on ecosystem.

In the frame of project PARC (work package 5 (WP5)) and more specifically in the frame of internal project "BPA alternatives and associated mixtures (Data gaps and NAM development)" we will focus on the investigation of potential adverse effects of certain individual BPA alternatives and "real-life" mixtures on different species to fill the
identified gaps regarding the potential harmful effects of bisphenols in the environment (activity of sub-task 5.1.2.). In addition, within this project, new technologies will be developed, such as *in silico* and modeling tools, read-across methodologies, and other advanced new approach methodologies (NAMs) covering both invertebrate and vertebrate organisms and several endpoints such as endocrine disruption (including metabolic disruption), non-genotoxic carcinogenicity, (developmental) neurotoxicity, immunotoxicity (activity of sub-task 5.2.2.).

In PARC, complementary to environmental related studies, BPA alternatives are also investigated in more human-oriented *in vitro* studies. The goal is to assess the hazards associated with these alternatives, contributing to their regulation as part of sub-task 5.1.1. This project aims to produce a complete dataset on the immunotoxicity, endocrine disruption activity, neurodevelopmental toxicity, genotoxicity, and non-genotoxic carcinogenicity of BPA alternatives and to explore mechanisms underlying the effects observed that would feed into adverse outcome pathways. The results of these projects, as well as the results of other projects within PARC including monitoring activities of BPA alternatives, can be exploited by decision makers and risk managers at European level. Decision makers bear the responsibility of action based on existing research data. Their key responsibility is to prioritize environmental and public health protection by implementing appropriate measures to regulate or restrict the use of identified hazardous BPA alternatives. Proactive measures grounded in current knowledge can effectively mitigate potential risks, ensuring the well-being of ecosystems and living organisms.

### CRediT authorship contribution statement

**Ondrej Adamovsky:** Writing – review & editing, Writing – original draft, Visualization, Supervision, Methodology, Funding acquisition, Conceptualization.
**Ksenia J. Groh:** Writing – review & editing, Writing – original draft.
**Anna Białk-Bielińska:** Writing – review & editing, Writing – original draft.
**Beate I. Escher:** Writing – review & editing, Writing – original draft.
**R. Beaudouin:** Writing – review & editing, Writing – original draft.
**L. Mora Lagares:** Writing – review & editing, Writing – original draft.
**K.E. Tollefsen:** Writing – review & editing, Writing – original draft.
**Martina Fenske:** Writing – review & editing, Writing – original draft.
**Ewa Mulkiewicz:** Writing – review & editing, Writing – original draft.
**Nicolas Creusot:** Writing – review & editing, Writing – original draft.
**Anita Sosnowska:** Writing – review & editing, Writing – original draft.
**Susana Loureiro:** Writing – review & editing, Writing – original draft.
**Jonny Beyer:** Writing – review & editing, Writing – original draft.
**Guillermo Repetto:** Writing – review & editing, Writing – original draft.
**Alja Štern:** Writing – review & editing, Writing – original draft.
**Isabel Lopes:** Writing – review & editing, Writing – original draft.
**Marta Monteiro:** Writing – review & editing, Writing – original draft.
**Andrea Zikova-Kloas:** Writing – review & editing, Writing – original draft.
**Tina Elersek:** Writing – review & editing, Writing – original draft.
**Marjan Vracko:** Writing – review & editing, Writing – original draft.
**Szymon Zdybel:** Writing – review & editing, Writing – original draft.
**Tomasz Puzyn:** Writing – review & editing, Writing – original draft.
**Weronika Koczur:** Writing – review & editing, Writing – original draft.
**Jane Ebsen Morthorst:** Writing – review & editing, Writing – original draft.
**Henrik Holbech:** Writing – review & editing, Writing – original draft.
**Gunnar Carlsson:** Writing – review & editing, Writing – original draft.
**Stefan Örn:** Writing – review & editing, Writing – original draft.
**Óscar Herrero:** Writing – review & editing, Writing – original draft.
**Ayesha Siddique:** Writing – review & editing, Writing – original draft.
**Matthias Liess:** Writing – review & editing, Writing – original draft.
**Georg Braun:** Writing – review & editing, Writing – original draft.
**Vanessa Srebny:** Writing – review & editing, Writing – original draft.
**Bojana Žegura:** Writing – review & editing, Writing – original draft.
**Nathalie Hinfray:** Writing – review & editing, Writing – original draft.
**François Brion:** Writing – review & editing, Writing – original draft.

**Dries Knapen:** Writing – review & editing, Writing – original draft.
**Ellen Vandeputte:** Writing – review & editing, Writing – original draft.
**Evelyn Stinckens:** Writing – review & editing, Writing – original draft.
**Lucia Vergauwen:** Writing – review & editing, Writing – original draft.
**Lars Behrendt:** Writing – review & editing, Writing – original draft.
**Maria Joao Silva:** Writing – review & editing, Writing – original draft.
**Ludek Blaha:** Funding acquisition.
**Katerina Kyriakopoulou:** Writing – original draft, Supervision.

### Declaration of competing interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

### Data availability

Data will be made available on request.

### Acknowledgment

This work was carried out in the framework of the European Partnership for the Assessment of Risks from Chemicals (PARC) and has received funding from the European Union's Horizon Europe research and innovation programme under Grant Agreement No 101057014. This work was also supported from the European Union's Horizon 2020 research and innovation program under grant agreement No 857560. Views and opinions expressed are however those of the author(s) only and do not necessarily reflect those of the European Union or the Health and Digital Executive Agency. Neither the European Union nor the granting authority can be held responsible for them. Ondrej Adamovsky and Ludek Blaha thank the RECETOX Research Infrastructure (No LM2023069) financed by the Ministry of Education, Youth and Sports, for supportive background. Dr. Adamovsky thanks also The Czech Science Foundation (GACR), project 23-13617L, for support in plastic and plastic additives research. Alja Štern, Bojana Žegura and Tina Elersek thank the ARIS program P1-0245 for supportive background. Susana Loureiro, Isabel Lopes and Marta S Monteiro thank the Foundation for Science and Technology and Ministry of Science, Technology and Higher Education (FCT/MCTES) for the financial support to CESAM (UIDP/50017/2020+UIDB/50017/2020+LA/P/0094/2020). Knut Erik Tolløsen and Jonny Beyer acknowledge the contribution from the Research Council of Norway projects 223268, 315969 and 342628. Dries Knapen, Lucia Vergauwen, Evelyn Stinckens and Ellen Vandeputte thank the University of Antwerp Research Fund (project 44602 and GOA project FFB180348/36572) and FPS Health, Food Chain Safety and Environment, DG Environment, in the framework of the National Action Plan on Endocrine Disruptors (NAPED, DGEM/DPPC/MJE/23079) for co-funding the work. Martina Fenske thanks the Federal Ministry for the Environment, Nature Conservation, Nuclear Safety and Consumer Protection (BMUV) for financial support (Action programme "Water management" of BfG and BMUV, Action 111

## Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.envint.2024.108728.

### References

Adamovsky, O., Buerger, A.N., Wormington, A.M., Ector, N., Griffitt, R.J., Bisesi, J.H., Martyniuk, C.J., 2018. The gut microbiome and aquatic toxicology: An emerging concept for environmental health. *Environ. Toxicol. Chem.* 37, 2758–2775. https://doi.org/10.1002/etc.4249.

O. Adamovsky, D. Basili, J. Bisesi, C. Martyniuk, S. Smatana, Adverse Outcomes and the Microbiome: A Guide to Comprehensive Characterization, Tech. Issue Pap. (2020). https://www.setac.org/general/custom.asp?page=SETACTechPapers.
identified gaps regarding the potential harmful effects of bisphenols in the environment (activity of sub-task 5.1.2.). In addition, within this project, new technologies will be developed, such as *in silico* and modeling tools, read-across methodologies, and other advanced new approach methodologies (NAMs) covering both invertebrate and vertebrate organisms and several endpoints such as endocrine disruption (including metabolic disruption), non-genotoxic carcinogenicity, (developmental) neurotoxicity, immunotoxicity (activity of sub-task 5.2.2.).

In PARC, complementary to environmental related studies, BPA alternatives are also investigated in more human-oriented *in vitro* studies. The goal is to assess the hazards associated with these alternatives, contributing to their regulation as part of sub-task 5.1.1. This project aims to produce a complete dataset on the immunotoxicity, endocrine disruption activity, neurodevelopmental toxicity, genotoxicity, and non-genotoxic carcinogenicity of BPA alternatives and to explore mechanisms underlying the effects observed that would feed into adverse outcome pathways. The results of these projects, as well as the results of other projects within PARC including monitoring activities of BPA alternatives, can be exploited by decision makers and risk managers at European level. Decision makers bear the responsibility of action based on existing research data. Their key responsibility is to prioritize environmental and public health protection by implementing appropriate measures to regulate or restrict the use of identified hazardous BPA alternatives. Proactive measures grounded in current knowledge can effectively mitigate potential risks, ensuring the well-being of ecosystems and living organisms.

### CRediT authorship contribution statement

**Ondrej Adamovsky:** Writing – review & editing, Writing – original draft, Visualization, Supervision, Methodology, Funding acquisition, Conceptualization.
**Ksenia J. Groh:** Writing – review & editing, Writing – original draft.
**Anna Białk-Bielińska:** Writing – review & editing, Writing – original draft.
**Beate I. Escher:** Writing – review & editing, Writing – original draft.
**R. Beaudouin:** Writing – review & editing, Writing – original draft.
**L. Mora Lagares:** Writing – review & editing, Writing – original draft.
**K.E. Tollefsen:** Writing – review & editing, Writing – original draft.
**Martina Fenske:** Writing – review & editing, Writing – original draft.
**Ewa Mulkiewicz:** Writing – review & editing, Writing – original draft.
**Nicolas Creusot:** Writing – review & editing, Writing – original draft.
**Anita Sosnowska:** Writing – review & editing, Writing – original draft.
**Susana Loureiro:** Writing – review & editing, Writing – original draft.
**Jonny Beyer:** Writing – review & editing, Writing – original draft.
**Guillermo Repetto:** Writing – review & editing, Writing – original draft.
**Alja Štern:** Writing – review & editing, Writing – original draft.
**Isabel Lopes:** Writing – review & editing, Writing – original draft.
**Marta Monteiro:** Writing – review & editing, Writing – original draft.
**Andrea Zikova-Kloas:** Writing – review & editing, Writing – original draft.
**Tina Elersek:** Writing – review & editing, Writing – original draft.
**Marjan Vracko:** Writing – review & editing, Writing – original draft.
**Szymon Zdybel:** Writing – review & editing, Writing – original draft.
**Tomasz Puzyn:** Writing – review & editing, Writing – original draft.
**Weronika Koczur:** Writing – review & editing, Writing – original draft.
**Jane Ebsen Morthorst:** Writing – review & editing, Writing – original draft.
**Henrik Holbech:** Writing – review & editing, Writing – original draft.
**Gunnar Carlsson:** Writing – review & editing, Writing – original draft.
**Stefan Örn:** Writing – review & editing, Writing – original draft.
**Óscar Herrero:** Writing – review & editing, Writing – original draft.
**Ayesha Siddique:** Writing – review & editing, Writing – original draft.
**Matthias Liess:** Writing – review & editing, Writing – original draft.
**Georg Braun:** Writing – review & editing, Writing – original draft.
**Vanessa Srebny:** Writing – review & editing, Writing – original draft.
**Bojana Žegura:** Writing – review & editing, Writing – original draft.
**Nathalie Hinfray:** Writing – review & editing, Writing – original draft.
**François Brion:** Writing – review & editing, Writing – original draft.

**Dries Knapen:** Writing – review & editing, Writing – original draft.
**Ellen Vandeputte:** Writing – review & editing, Writing – original draft.
**Evelyn Stinckens:** Writing – review & editing, Writing – original draft.
**Lucia Vergauwen:** Writing – review & editing, Writing – original draft.
**Lars Behrendt:** Writing – review & editing, Writing – original draft.
**Maria Joao Silva:** Writing – review & editing, Writing – original draft.
**Ludek Blaha:** Funding acquisition.
**Katerina Kyriakopoulou:** Writing – original draft, Supervision.

### Declaration of competing interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

### Data availability

Data will be made available on request.

### Acknowledgment

This work was carried out in the framework of the European Partnership for the Assessment of Risks from Chemicals (PARC) and has received funding from the European Union's Horizon Europe research and innovation programme under Grant Agreement No 101057014. This work was also supported from the European Union's Horizon 2020 research and innovation program under grant agreement No 857560. Views and opinions expressed are however those of the author(s) only and do not necessarily reflect those of the European Union or the Health and Digital Executive Agency. Neither the European Union nor the granting authority can be held responsible for them. Ondrej Adamovsky and Ludek Blaha thank the RECETOX Research Infrastructure (No LM2023069) financed by the Ministry of Education, Youth and Sports, for supportive background. Dr. Adamovsky thanks also The Czech Science Foundation (GACR), project 23-13617L, for support in plastic and plastic additives research. Alja Štern, Bojana Žegura and Tina Elersek thank the ARIS program P1-0245 for supportive background. Susana Loureiro, Isabel Lopes and Marta S Monteiro thank the Foundation for Science and Technology and Ministry of Science, Technology and Higher Education (FCT/MCTES) for the financial support to CESAM (UIDP/50017/2020+UIDB/50017/2020+LA/P/0094/2020). Knut Erik Tollefsen and Jonny Beyer acknowledge the contribution from the Research Council of Norway projects 223268, 315969 and 342628. Dries Knapen, Lucia Vergauwen, Evelyn Stinckens and Ellen Vandeputte thank the University of Antwerp Research Fund (project 44602 and GOA project FFB180348/36572) and FPS Health, Food Chain Safety and Environment, DG Environment, in the framework of the National Action Plan on Endocrine Disruptors (NAPED, DGEM/DPPC/MJE/23079) for co-funding the work. Martina Fenske thanks the Federal Ministry for the Environment, Nature Conservation, Nuclear Safety and Consumer Protection (BMUV) for financial support (Action programme "Water management" of BfG and BMUV, Action 111

## Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.envint.2024.108728.

### References

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O. Adamovsky, D. Basili, J. Bisesi, C. Martyniuk, S. Smatana, Adverse Outcomes and the Microbiome: A Guide to Comprehensive Characterization, Tech. Issue Pap. (2020). https://www.setac.org/general/custom.asp?page=SETACTechPapers.
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