# A critical review to identify data gaps and improve risk assessment of bisphenol A alternatives for human health

Sakina Mhaouty-Kodja, Daniel Zalko, Sabrina Tait, Emanuela Testai, Catherine Viguié, Emanuela Corsini, Nathalie Grova, Franca Maria Buratti, Nicolas Cabaton, Lucia Coppola, et al.

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## ► To cite this version:

Sakina Mhaouty-Kodja, Daniel Zalko, Sabrina Tait, Emanuela Testai, Catherine Viguié, et al. A critical review to identify data gaps and improve risk assessment of bisphenol A alternatives for human health. Critical Reviews in Toxicology, 2024, pp.1-58. 10.1080/10408444.2024.2388712. hal-04749021

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# A critical review to identify data gaps and improve risk assessment of bisphenol A alternatives for human health

Sakina Mhaouty-Kodja, Daniel Zalko, Sabrina Tait, Emanuela Testai, Catherine Viguié, Emanuela Corsini, Nathalie Grova, Franca Maria Buratti, Nicolas J. Cabaton, Lucia Coppola, Antonio De la Vieja, Maria Dusinska, Naouale El Yamani, Valentina Galbiati, Patricia Iglesias-Hernández, Yvonne Kohl, Ambra Maddalon, Francesca Marcon, Lydie Naulé, Elise Rundén-Pran, Francesca Salani, Nicoletta Santori, Mónica Torres-Ruiz, Jonathan D. Turner, Ondrej Adamovsky, Kiara Aiello-Holden, Hubert Dirven, Henriqueta Louro & Maria João Silva

**To cite this article:** Sakina Mhaouty-Kodja, Daniel Zalko, Sabrina Tait, Emanuela Testai, Catherine Viguié, Emanuela Corsini, Nathalie Grova, Franca Maria Buratti, Nicolas J. Cabaton, Lucia Coppola, Antonio De la Vieja, Maria Dusinska, Naouale El Yamani, Valentina Galbiati, Patricia Iglesias-Hernández, Yvonne Kohl, Ambra Maddalon, Francesca Marcon, Lydie Naulé, Elise Rundén-Pran, Francesca Salani, Nicoletta Santori, Mónica Torres-Ruiz, Jonathan D. Turner, Ondrej Adamovsky, Kiara Aiello-Holden, Hubert Dirven, Henriqueta Louro & Maria João Silva (22 Oct 2024): A critical review to identify data gaps and improve risk assessment of bisphenol A alternatives for human health, Critical Reviews in Toxicology, DOI: 10.1080/10408444.2024.2388712

**To link to this article:** https://doi.org/10.1080/10408444.2024.2388712

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© 2024 The Author(s). Published by Informa UK Limited, trading as Taylor & Francis Group

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Published online: 22 Oct 2024.

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# A critical review to identify data gaps and improve risk assessment of bisphenol A alternatives for human health

Sakina Mhaouty-Kodja<sup>a*</sup>, Daniel Zalko<sup>b*</sup>, Sabrina Tait<sup>c*</sup>, Emanuela Testai<sup>d*</sup> (✉), Catherine Viguié<sup>b*</sup>, Emanuela Corsini<sup>e</sup>, Nathalie Grova<sup>f</sup>, Franca Maria Buratti<sup>d</sup>, Nicolas J. Cabaton<sup>b</sup>, Lucia Coppola<sup>c</sup>, Antonio De la Vieja<sup>g</sup>, Maria Dusinska<sup>h</sup>, Naouale El Yamani<sup>h</sup>, Valentina Galbiati<sup>e</sup>, Patricia Iglesias-Hernández<sup>g</sup>, Yvonne Kohl<sup>i</sup>, Ambra Maddalon<sup>e</sup>, Francesca Marcon<sup>d</sup>, Lydie Naulé<sup>a</sup>, Elise Rundén-Pran<sup>h</sup>, Francesca Salani<sup>d</sup>, Nicoletta Santori<sup>d</sup>, Mónica Torres-Ruiz<sup>j</sup>, Jonathan D. Turner<sup>f</sup>, Ondrej Adamovsky<sup>k</sup>, Kiara Aiello-Holden<sup>l</sup>, Hubert Dirven<sup>m</sup>, Henriqueta Louro<sup>n,o</sup> and Maria João Silva<sup>n,o</sup>

<sup>a</sup>CNRS UMR 8246, INSERM U1130, Neuroscience Paris Seine – Institut de Biologie Paris Seine, Sorbonne Université, Paris, France; <sup>b</sup>INRAE, UMR1331 Toxalim (Research Center in Food Toxicology), Université de Toulouse, INRAE, ENVT, INP-Purpan, UT3, Toulouse, France; <sup>c</sup>Center for Gender-Specific Medicine, Istituto Superiore di Sanità, Rome, Italy; <sup>d</sup>Department of Environment and Health, Mechanisms, Biomarkers and Models Unit, Istituto Superiore di Sanità, Rome, Italy; <sup>e</sup>Department of Pharmacological and Biomolecular Sciences ‘Rodolfo Paoletti’, Università degli Studi di Milano – School of Pharmacy, Milan, Italy; <sup>f</sup>Department of Infection and Immunity, Immune Endocrine Epigenetics Research Group, Luxembourg Institute of Health, Esch-Sur-Alzette, Luxembourg; <sup>g</sup>Endocrine Tumor Unit from Chronic Disease Program (UFIEC), Instituto de Salud Carlos III (ISCIII), Majadahonda, Madrid, Spain; <sup>h</sup>Department for Environmental Chemistry, Health Effects Laboratory, NILU-Norwegian Institute for Air Research, Kjeller, Norway; <sup>i</sup>Fraunhofer Institute for Biomedical Engineering IBMT, Sulzbach, Germany; <sup>j</sup>National Center for Environmental Health (CNSA), Instituto de Salud Carlos III (ISCIII), Majadahonda, Madrid, Spain; <sup>k</sup>Faculty of Science, Masaryk University, RECETOX, Brno, Czech Republic; <sup>l</sup>German Federal Institute for Risk Assessment (BfR), Berlin, Germany; <sup>m</sup>Department of Chemical Toxicology – Division of Climate and the Environment, Norwegian Institute of Public Health, Oslo, Norway; <sup>n</sup>Department of Human Genetics, National Institute of Health Dr. Ricardo Jorge, Lisbon, Portugal; <sup>o</sup>Centre for Toxicogenomics and Human Health, Nova Medical School, Universidade Nova de Lisboa, Lisbon, Portugal

## ABSTRACT

Bisphenol A (BPA), a synthetic chemical widely used in the production of polycarbonate plastic and epoxy resins, has been associated with a variety of adverse effects in humans including metabolic, immunological, reproductive, and neurodevelopmental effects, raising concern about its health impact. In the EU, it has been classified as toxic to reproduction and as an endocrine disruptor and was thus included in the candidate list of substances of very high concern (SVHC). On this basis, its use has been banned or restricted in some products. As a consequence, industries turned to bisphenol alternatives, such as bisphenol S (BPS) and bisphenol F (BPF), which are now found in various consumer products, as well as in human matrices at a global scale. However, due to their toxicity, these two bisphenols are in the process of being regulated. Other BPA alternatives, whose potential toxicity remains largely unknown due to a knowledge gap, have also started to be used in manufacturing processes. The gradual restriction of the use of BPA underscores the importance of understanding the potential risks associated with its alternatives to avoid regrettable substitutions. This review aims to summarize the current knowledge on the potential hazards related to BPA alternatives prioritized by European Regulatory Agencies based on their regulatory relevance and selected to be studied under the European Partnership for the Assessment of Risks from Chemicals (PARC): BPE, BPAP, BPP, BPZ, BPS-MAE, and TCBPA. The focus is on data related to toxicokinetic, endocrine disruption, immunotoxicity, developmental neurotoxicity, and genotoxicity/carcinogenicity, which were considered the most relevant endpoints to assess the hazard related to those substances. The goal here is to identify the data gaps in BPA alternatives toxicology and hence formulate the future directions that will be taken in the frame of the PARC project, which seeks also to enhance chemical risk assessment methodologies using new approach methodologies (NAMs).

**Abbreviations:** ADME: absorption, distribution, metabolism, and excretion; AhR: aryl hydrocarbon receptor; AR: androgen receptor; AUC: area under the curve; BADGE: bisphenol A diglycidyl ether; BBB: blood-brain-barrier; BCRP: breast cancer resistance protein; BDNF: brain-derived neurotrophic factor; BFDGE: bisphenol F diglycidyl-ether; BPA: bisphenol A; BPA-G: BPA mono-glucuronide; BPAF: bisphenol AF; BPAP: bisphenol AP; BPB: bisphenol B; BPBP: bisphenol BP; BPC: bisphenol C; BPE: bisphenol E; BPF: bisphenol F; BPFDGE: bisphenol F diglycidylether; BPFL: bisphenol FL; BPM: bisphenol M; BPP: bisphenol

## ARTICLE HISTORY

Received 1 April 2024
Revised 29 July 2024
Accepted 30 July 2024

## KEYWORDS

Bisphenol A alternatives;
endocrine disruption;
metabolism; toxicokinetic;
immunotoxicity; develop-
mental neurotoxicity;
genotoxicity; carcinogenesis;
human health

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<footnote>*These authors contributed equally to this work.</footnote>
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P; BPPH: bisphenol PH; BPS: bisphenol S; BPS-G: BPS-glucuronide; BPS-IP: BPS-isopropyl; BPS-MAE: bisphenol S 4-allyl ether; BPZ: bisphenol Z; CAR: constitutive androstane receptor; CGCs: cerebellar granule cells; CLP: Classification, Labelling, and Packaging; CHO: Chinese hamster ovary; COMT: catechol O methyl transferase; CTA: cell transformation assay; CYP: cytochrome P450; DHEA: dehydroepiandrosterone; DHT: 5α-dihydrotestosterone; DIO: iodothyronine deiodinase; DNT: developmental neurotoxicity; E2: 17β-estradiol; EC: effective concentration; ECHA: European Chemicals Agency; ED: endocrine disrupting; EDC: endocrine disrupting chemical; EFSA: European Food Safety Authority; ER: estrogen receptor; ERR: estrogen-related receptor; FOXE1: forkhead box protein E1; FXR: farnesoid X receptor; γH2AX: histone family member X; GIVIMP: Guidance Document on Good In Vitro Method Practices; GM-CSF: granulocyte-macrophage colony-stimulating factor; GPER: G protein-coupled estrogen receptor; GR: glucocorticoid receptor; hCG: human chorionic gonadotropin; HED: human equivalent dose; hESCs: human embryonic stem cells; HFR: halogenated flame retardants; HIM: human intestine microsomes; HLC: human liver cytosol; HLM: human liver microsomes; HSA: human serum albumin; HSD: hydroxysteroid dehydrogenase; HTTK: high throughput toxicokinetics; IFN-γ: interferon gamma; IL: interleukin; Il1rap1: interleukin 1 receptor accessory protein like 1; NR: nuclear receptor; LXR: liver X receptor; MCT8: monocarboxylate transporter 8; MDC: metabolism disrupting chemical; MR: mineralocorticoid receptor; NAM: new approach methodology; Ncam2: neural cell adhesion molecule; NCoR: nuclear receptor corepressor; NF-κB: nuclear factor-kappaB; NGRA: new generation risk assessment; NIS: sodium/iodide symporter; NKX2.1: NK homeobox protein 1; NK: natural killer; NSCs: neural stem cells; OECD: Organisation for Economic Cooperation and Development; PARC: Partnership for the Assessment of Risks from Chemicals; PAX8: paired box 8; PBK: physiological based kinetic; PPAR: peroxisome proliferator-activated receptor; PR: progesterone receptor; PXR: pregnane X receptor; QSAR: quantitative structure-activity relationship; QLVIVE: quantitative in vitro to in vivo extrapolation; RAR: retinoic acid receptor; REACH: European regulation on the Registration, Evaluation, Authorisation and restriction of Chemicals; ROR: RAR-related orphan receptor; ROS: reactive oxygen species; RXR: retinoid X receptor; SCENIHR: Scientific Committee on Emerging and Newly Identified Health Risks; SERM: selective estrogen receptor modulator; SHE: Syrian hamster embryo; SMRT: silencing mediator for retinoid and thyroid-hormone receptor; SNAP: synaptosomal-associated protein; StAR: steroidogenic acute regulatory protein; SULT: sulfotransferase; SVHC: substance of very high concern; T3: triiodothyronine; T4: thyroxine; TBBPA: tetra-bromo bisphenol A; TBG: thyroxine-binding globulin; TCBPA: tetra-chloro bisphenol A; TDI: tolerable daily intake; TG: thyroglobulin; TGF: transforming growth factor; Th: T helper; TH: thyroid hormone; THR: thyroid hormone receptor; TK: toxicokinetics; TMBPA: tetra-methyl BPA; TNF-α: tumor necrosis factor alpha; TPO: thyroperoxidase; TSH: thyroid-stimulating hormone; TSHR: thyroid-stimulating hormone receptor; TTR: transthyretin; UGT: UDP glucuronosyltransferase; VDR: vitamin D receptor; WoE: weight of evidence; WHO: World Health Organization; XME: xenobiotic metabolizing enzymes

Table of contents

1. Introduction ... 3

2. ADME (absorption, distribution, metabolism, and excretion) processes of BPA alternatives ... 4

2.1. General considerations ... 4

2.2. Information available on BPA-alternatives kinetics ... 5

2.3. Placental and blood-brain-barrier (BBB) transfer ... 6

2.4. Conjugation pathways: a major metabolic reaction for BPA alternatives ... 6

2.5. Oxidative metabolism of BPA alternatives ... 7

2.6. Isoform specific metabolism of BPA analogues ... 7

2.7. Halogenated bisphenols and other specific alternatives ... 8

2.8. Concluding considerations on the kinetics of BPA alternatives ... 10

3. Endocrine disrupting activity of BPA alternatives ... 10

3.1. General considerations ... 10

3.2. Interaction and effects on nuclear steroid, PPARs, and other receptors ... 10

3.2.1. Estrogen receptors (ERα, ERβ, ERRγ, GPER) ... 10

3.2.2. Androgen receptor (AR) ... 11

3.2.3. Progesterone receptor (PR) ... 12

3.2.4. Glucocorticoid receptor (GR) ... 12

3.2.5. Peroxisome proliferator activated receptors (PPARα, β/δ, γ) ... 13

3.2.6. Other receptors ... 13

3.2.7. Concluding considerations ... 15

3.3. Effects on thyroid hormone system pathways ... 15

3.3.1. Effects on TH levels ... 15

3.3.2. Effects on thyroid hormone receptors (THRs) ... 15

3.3.3. Effects on TH biosynthesis ... 16

3.3.4. Interaction with TH serum transporters ... 17

3.3.5. Interaction with TH plasma membrane transporters ... 17

3.3.6. Concluding considerations ... 17

3.4. Effects on steroidogenesis ... 17

3.4.1. Concluding considerations ... 18

4. Immunotoxicity of BPA alternatives ... 18

4.1. General considerations ... 18

4.2. Information on the immunotoxicity of BPA alternatives ... 18

4.3. Concluding considerations ... 20

5. Developmental neurotoxicity of BPA alternatives ... 20

5.1. General considerations ... 20

5.2. Neurotoxic effects of BPA analogues and halogenated bisphenols ... 20

5.2.1. BPA analogues ... 20

5.2.2. Halogenated bisphenols ... 21

5.3. Concluding considerations ... 22
halogenation. TBBPA was identified by ECHA as a carcinogen and has been placed in the candidate list of SVHC (ECHA 2023). These substances have been used to manufacture flame proofed epoxy resins, with TBBPA being a major flame retardant in terms of tonnage/use. It cannot be ruled out that other halogenated BPA analogues or alternatives (e.g. mono- to tetra-chloro BPA) can form in the environment due to the reaction of BPA with chlorinated waters (Plattard et al. 2021). In addition, photodegradation products of BPA chlorinated derivatives have been identified in wastewater treatment plant effluent under sunlight irradiation (Wan et al. 2020). Human exposure to chlorinated analogues was clearly demonstrated when they were detected in adipose tissue from women living in Southeast Spain (Fernandez et al. 2007) and later confirmed in placenta samples (Jiménez-Díaz et al. 2010). Subsequently, their presence in human tissue has been confirmed in several other countries (Andra et al. 2015). Therefore, the occurrence of mixed brominated/chlorinated bisphenols cannot be excluded. For most of these bisphenol alternatives, information on biotransformation and toxicokinetics (TK), ED activity, developmental neurotoxicity (DNT), genotoxicity, and carcinogenicity are either absent or too limited for a full hazard assessment and characterization.

A major goal of the Partnership for the Assessment of Risks from Chemicals (PARC, https://www.eu-parc.eu) is to advance the knowledge in the risk assessment of chemicals by closing data gaps on hazards for some substances and developing and integrating new approach methodologies (NAMs) (Marx-Stoelting et al. 2023). To this aim, a variety of studies are dedicated to the hazard assessment of substances included in a previously defined priority list (PARC D2.1, available at https://www.eu-parc.eu/sites/default/files/2023-09/PARC_D2.1.pdf) using several methodological approaches. On the one hand, for selected substances, e.g. BPA alternatives and some natural toxins, the identified regulatory data gaps comprising immunotoxicity and endocrine disruption, among other effects, will be filled by using test methods recommended by International Organizations (e.g. Organisation for Economic Cooperation and Development (OECD) Test Guidelines) in order to facilitate the regulatory acceptance of the data generated. On the other hand, NAMs will be developed to contribute to decrease uncertainties and improve those substances risk assessment. Indeed, seven distinct, but interacting scientific priorities areas were proposed to reach the goal of implementing NAMs into the new generation risk assessment (NGRA) (Escher et al. 2022). Those priorities are: production of TK data, with a focus on physiological based kinetic (PBK) modeling, development of additional adverse outcome pathways (AOPs) or AOP networks, development of advanced cell culture models including organ on chips, expensive, human susceptibility, data integration, new concepts in risk assessment (with cross-sectional activities, acceptance and assessment criteria, to update the guidance on how to integrate NAMs in human risk assessment), generally relevant for many chemicals, including BPA alternatives.

Among the group of BPA alternatives, some substances were prioritized based on the existent knowledge and discussions with the ECHA and EFSA experts, the relevant European stakeholders interested in the safety evaluation of these substances. The existence of evidence of human exposure and

the fact that their hazard had not been assessed by industry were considered as criteria for the selection. The specific project on the hazard assessment of BPA alternatives aims to fill the existing data gaps by providing a complete data set on four endpoints prioritized within PARC for human health, including ED activity, immunotoxicity, DNT, and genotoxicity/carcinogenicity. In addition, data on the toxicokinetic and metabolism will be produced for some compounds. A tiered approach will be followed for the hazard assessment of the selected BPA alternatives, with the collection of the published information representing the first step. In the scope of PARC, another project is addressing the effects of the same BPA alternatives on the ecosystem. With the objective of identifying research gaps on the potential deleterious effects of those compounds on the environment health, a review was recently published that describes the current knowledge and approaches used to assess their toxicity in invertebrate and vertebrate models (Adamovsky et al. 2024). Therefore, data and information from those alternative animal models are out of the scope of the present review.

This review intends to critically address the current knowledge concerning the hazard of the BPA alternatives, particularly focusing on substances not currently subjected to any restriction and that were prioritized in PARC: BPE, BPAP, BPP, BPZ, bisphenol S 4-allyl ether (BPS-MAE), and TCBPA. Other alternative substances, such as Pergafast-201 and TBBPA, were included for specific endpoints (Table 1). Additionally, information related to BPA and its most commonly used analogues for which data are already available (e.g. BPS or BPF), is briefly reported for comparison. Besides offering a substantial update of the literature since the review by Pelch K et al. (2019), this review includes novel TK and metabolism information on the selected alternatives, drawing conclusions and identifying main gaps in knowledge about their hazard or risks. This information may guide the production of in vitro data that can be used as input for specific PBK modeling toward a NGRA approach, as foreseen by PARC, to avoid future regrettable substitutions.

## 2. ADME (absorption, distribution, metabolism, and excretion) processes of BPA alternatives

### 2.1. General considerations

The amount of any chemical that can reach and exert effects in target tissues leading to systemic toxicity is dependent on ADME processes. Already in 2010, the OECD adopted a guideline for the testing of chemicals' TK, which highlights the usefulness of TK data in many steps of the risk assessment of chemical substances (OECD 2010). Similar recommendations were reiterated over the years by the European Commission and other Agencies (European Commission 2012; EFSA 2014), placing TK among the key research priority topics, to accelerate the move toward the reduction of animal testing (3R concept) and promote a more mechanistic understanding of toxicology. More recently, EFSA pointed out that the production of TK data, with a focus on PBK modeling, is one among the seven distinct, but interacting scientific priorities areas proposed for reaching the goal of implementing NAMs into the NGRA (Escher et al. 2022).
<figure><img src="image_18.png" /><figcaption>**Table 1.** List of prioritized BPA alternatives including: CAS number, common and IUPAC name, chemical structure and type of central carbon (quaternary/not quaternary).</figcaption></figure>

<table><thead><tr><th>CAS Number</th><th>Common Name<br>IUPAC Name</th><th>Chemical structure and central carbon:<br>quaternary/not quaternary</th><th></th></tr></thead><tbody><tr><td>80-05-7</td><td>*BPA; Bisphenol A/<br>4-[2-(4-hydroxyphenyl)propan-2-yl]phenol</td><td><img src="image_0_0_0.png" alt="Chemical structure of 4-[2-(4-hydroxyphenyl)propan-2-yl]phenol"></td><td>●</td></tr><tr><td>80-09-1</td><td>*BPS; Bisphenol S/<br>4,4'-sulphonyldiphenol</td><td><img src="image_0_0_1.png" alt="Chemical structure of 4,4'-sulphonyldiphenol"></td><td>●</td></tr><tr><td>620-92-8</td><td>*BPF; Bisphenol F/<br>4,4'-methylenediphenol</td><td><img src="image_0_0_2.png" alt="Chemical structure of 4,4'-methylenediphenol"></td><td>●</td></tr><tr><td>843-55-0</td><td>BPZ; Bisphenol Z/<br>4-[1-(4-hydroxyphenyl)cyclohexyl]phenol</td><td><img src="image_0_0_3.png" alt="Chemical structure of 4-[1-(4-hydroxyphenyl)cyclohexyl]phenol"></td><td>●</td></tr><tr><td>2081-08-5</td><td>BPE; Bisphenol E/<br>4-[1-(4-hydroxyphenyl)ethyl]phenol</td><td><img src="image_0_0_4.png" alt="Chemical structure of 4-[1-(4-hydroxyphenyl)ethyl]phenol"></td><td>●</td></tr><tr><td>97042-18-7</td><td>BPS-MAE/<br>4-(4-prop-2-enoxyphenyl)sulfonylphenol</td><td><img src="image_0_0_5.png" alt="Chemical structure of 4-(4-prop-2-enoxyphenyl)sulfonylphenol"></td><td>●</td></tr><tr><td>232938-43-1</td><td>Pergafast 201/<br>[3-[(4-methylphenyl) sulfonylcarbamoylamino]<br>phenyl] 4-methylbenzenesulfonate</td><td><img src="image_0_0_6.png" alt="Chemical structure of 3-[(4-methylphenyl) sulfonylcarbamoylamino] phenyl 4-methylbenzenesulfonate"></td><td>●</td></tr><tr><td>2167-51-3</td><td>BPP; Bisphenol P/<br>4-[1-(4-hydroxyphenyl)-1-phenylethyl]phenol</td><td><img src="image_0_0_7.png" alt="Chemical structure of 4-[1-(4-hydroxyphenyl)-1-phenylethyl]phenol"></td><td>●</td></tr><tr><td>1571-75-1</td><td>BPAP; Bisphenol AP/<br>4-[1-(4-hydroxyphenyl)-1-phenylethyl]phenol</td><td><img src="image_0_0_8.png" alt="Chemical structure of 4-[1-(4-hydroxyphenyl)-1-phenylethyl]phenol"></td><td>●</td></tr><tr><td>79-95-8</td><td>TCBPA; tetrachlorobisphenol A/<br>2,6-dichloro-4-[2-(3,5-dichloro-4-hydroxyphenyl)propan-2-yl]phenol</td><td><img src="image_0_0_9.png" alt="Chemical structure of 2,6-dichloro-4-[2-(3,5-dichloro-4-hydroxyphenyl)propan-2-yl]phenol"></td><td>●</td></tr><tr><td>79-94-7</td><td>TBBPA; tetrabromobisphenol A/<br>2,6-dibromo-4-[2-(3,5-dibromo-4-hydroxyphenyl)propan-2-yl]phenol</td><td><img src="image_0_0_10.png" alt="Chemical structure of 2,6-dibromo-4-[2-(3,5-dibromo-4-hydroxyphenyl)propan-2-yl]phenol"></td><td>●</td></tr></tbody></table>

Over the last decades, the increasing knowledge about the ADME characteristics of BPA has been extremely useful to decrease uncertainties associated with the derivation of health-based values. Indeed, TK knowledge allowed that species and age-related differences in BPA systemic bioavailability can be considered, and these studies also contributed to the development of adequate PBK modeling to derive a human equivalent dose (HED) (EFSA 2010, 2015, 2023). The considerable data available as regards BPA TK has been recently used as a case study to show

how kinetics data, integrated with a computational workflow, can facilitate quantitative *in vitro* to *in vivo* extrapolation (QIVIVE) (Loizou et al. 2021).

## 2.2. Information available on BPA-alternatives kinetics

As regards BPA alternatives, data are available only for a limited number of substances. When useful, these have been

<footnote>*BPA and its analogues BPS and BPF, all data reach substances, are used in this review for comparison (substances in grey). Further details on the prioritization process can be found at: https://www.eu-parc.eu/sites/default/files/2023-08/PARC_AD5.1.pdf.</footnote>
<figure><img src="image_19.png" /></figure>

briefly cited here for comparison. The best documented compound is BPS, which, for this reason, has not been included in the priority list of substances considered within the PARC project. A review has recently been published (Beausoleil et al. 2022) indicating that the estimated BPS human plasma clearance (0.92 L/min) is lower than that of BPA, suggesting a less efficient elimination (Grandin F et al. 2017; Gayrard et al. 2020) and potentially increased toxicity (e.g. on reproductive system) compared to BPA (Beausoleil et al. 2022). Studies comparing BPA and BPS ADME data following oral or dermal exposure, have clearly shown that, despite the similarities between the two molecules, significant TK differences exist (Thayer et al. 2015; Karrer et al. 2018; Oh et al. 2018; Waidyanatha et al. 2018; Liu J and Martin 2019; Champmartin et al. 2020; EFSA 2020; Khmiri et al. 2020; Reale et al. 2020). These observations underline the need to acquire ADME data for all BPA alternatives. In addition, species differences in TK have been reported (Sonker et al. 2021), which should be considered in the risk assessment process, when rodent data are used, as already achieved for BPA with the derivation of the HED. Three TK models have been proposed for BPS (Karrer et al. 2018; Oh et al. 2018; Gingrich et al. 2021). The model of Oh et al. (2018) has the advantage of being based on human data, but it is not a physiologically based model. The two other models are PBK but are based on quantitative structure-activity relationship (QSAR) and/or animal data (Corbel et al. 2013; Grandin FC et al. 2018; Gingrich et al. 2021). More recently, a model of human dermal absorption of BPA, BPS, BPF, and bisphenol AF (BPAF) was published (Hu M et al. 2023), in which a parallel-layered skin compartment model was integrated into the existing PBK models for BPS (Karrer et al. 2018). It can be derived that, after dermal exposure, women experienced an internal exposure to BPS characterized by an area under the curve (AUC, i.e. the definite integral of the concentration of a chemical in blood plasma as a function of time) and $C_{max}$ (the highest concentration in blood) values 11-fold lower when compared to BPA. For BPF and BPAF, the same parameters were 3- and 40-fold lower than for BPA, respectively (Hu M et al. 2023).

Another recent study investigated the TK behaviors of mixtures of 11 BPA alternatives (among which BPS, BPAF, BPB, BPF, BPZ, BPAP, BPP, BPM, and BPFL) following intravenous or oral administration to piglets (Gély et al. 2023). The systemic concentration reached by each substance was compared with that of BPA using the TK predicted by the generic High Throughput Toxicokinetics (HTTK) model package. The findings indicated that, at the same oral dose, all the tested BPA analogues produced a higher systemic exposure (measured as AUC) than BPA (2- to 4-fold for BPAF, BPB, and BPZ; 7- to 20-fold for BPAP, BPP, and BPF and 150-fold for BPS). The observed differences were attributed to variations in the systemic bioavailability of the substances. The results also indicated that conversely to other bisphenols, BPP, bisphenol M (BPM), and bisphenol FL (BPFL) showed a low urinary excretion, an important aspect to consider in predicting human exposure based on urine biomonitoring (Gély et al. 2023).

## 2.3. Placental and blood-brain-barrier (BBB) transfer

Some publications have documented the transfer of BPA and its alternatives through the placental barrier and the BBB. Using an *ex vivo* human placental perfusion model (Gély et al. 2021), bisphenols were shown to greatly differ in placental transport efficiency, with BPAP, BPE, BPF, BPB, and BPA being transferred by passive diffusion. However, BPZ, BPP, and BPAF, among others, had a lower transfer, suggesting weak diffusional permeability and/or the action of efflux protein. The placental transfer rate of BPS was also very limited (Gély et al. 2021): only 0.40% of the maternal BPS dose was transferred, 10 times lower than that of BPA (Grandin FC et al. 2019).

Differences in TK among 15 bisphenols (BPA, BPF, BPAP, BPP, BPZ, BPE, BPS, BPAF, and among others) have also been demonstrated in humans by studying 60 triplets of maternal plasma, cord plasma, and placenta samples from pregnant women in South China (Pan et al. 2020). Besides BPA, also BPS, BPAF and BPE were measured in all maternal plasma samples and detected with high frequency (from 80 to 100%) in mother-child pairs. Significant correlations between maternal plasma and cord plasma concentrations were observed for the most frequently detected bisphenols. The comparison between maternal and cord blood levels of bisphenols showed a different behavior regarding placental transfer efficiencies, with BPAF showing the highest efficiency and suggesting potential accumulation in the fetus (Pan et al. 2020).

Information about the transport across the BBB is extremely limited and no published data have been found for the BPA alternatives targeted in this review. However, docking modeling suggests that, based on their general structure, bisphenols are expected to bind to the same cavity of the breast cancer resistance protein (BCRP) efflux transporter, similarly to BPA and other inhibitors, in human *in vitro* BBB models (Nickel and Mahringer 2014; Engdahl et al. 2021).

## 2.4. Conjugation pathways: a major metabolic reaction for BPA alternatives

Based on the large database on BPA TK, it has been assumed that structurally related alternatives readily undergo conjugation to corresponding glucuronides and/or sulfates. Indeed, given the phenolic nature of these substances, no first-step oxidation is required prior to conjugation. The occurrence of conjugation pathways was demonstrated for BPA and several bisphenols (Knaak and Sullivan 1966). BPA mono-glucuronide (BPA-G) is the major metabolite of BPA in rodents (Snyder et al. 2000; Zalko et al. 2003) as well as in humans and other primates (Völkel et al. 2002; VandeVoort et al. 2016). This pathway was also extensively documented for BPS *in vitro* and *in vivo*, including in humans (Gys et al. 2018; Waidyanatha et al. 2018), as well as for some other alternatives, including BPF (Cabaton et al. 2006, 2008), BPAF (Li M et al. 2013), and TBBPA (Fini et al. 2012). In rodents, BPF was found to be efficiently metabolized into a sulfate conjugate following oral administration (Cabaton et al. 2006), but for most analogues, the relative contribution of glucuronidation vs. sulfation was not fully characterized. Notably, the
<figure><img src="image_20.png" /></figure>

occurrence of doubly conjugated metabolites (glucuronide and/or sulfate) was also demonstrated. Conversely to the parent BPA molecule, BPA-G was shown not to be an activator of nuclear ERs (Matthews et al. 2001). However, some reports have suggested that BPA-G has biological effects in vitro (Viñas et al. 2013; Boucher et al. 2015), but these should be considered with much caution, due to the high BPA-G concentrations tested and the possible back conversion of BPA-G into its unconjugated form that has been demonstrated in vivo (Corbel et al. 2015; Gauderat et al. 2016). Consequently, BPA glucuronidation is generally regarded as a detoxification pathway and, likewise, the sulfation pathway, despite the lack of conclusive published data. Of note, some conflicting results exist about the ER activities of BPS-sulfate (Kang et al. 2014; Skledar et al. 2016). With the exception of the data reported for few BPA alternatives, very limited information exists on the metabolic pathways followed by the substances selected to be studied under the PARC project.

## 2.5. Oxidative metabolism of BPA alternatives

The occurrence of oxidations, driven by cytochrome P450 (CYP) activity, has also been clearly demonstrated for BPA, as well as for few other alternatives. These pathways are still very much unexplored and warrant further investigations based on chemical similarities with BPA. Indeed, bisphenols oxidation may trigger the production of reactive intermediates and metabolites, which may explain some observed reactive oxygen species (ROS) related effects, such as immunotoxic and genotoxic effects (Zalko et al. 2006). Several (and distinct) oxidative metabolic pathways have been identified based on BPA studies. Their occurrence is closely conditioned by the chemistry of the molecules, which brings useful clues for further investigations to be carried out on not yet studied alternatives.

First, oxidation can occur ortho to one of the hydroxy functions, leading to the formation of a reactive catechol, and eventually of the corresponding quinone (Jaeg et al. 2004; Schmidt et al. 2013). Since these two metabolites may form DNA adducts (Atkinson and Roy 1995), this specific phase I reaction can be considered as a bioactivation route, unless BPA-catechol is detoxified by catechol-O-methyltransferase (COMT) activities (Zalko et al. 2003). Although catechols and related chemically reactive intermediates are quantitatively minor metabolites, few studies have demonstrated the occurrence of this pathway for BPS, BPF, BPAF, BPZ, BPB, and BPC (Yoshihara et al. 2004; Schmidt et al. 2013). Of note, BPA analogues bearing bulky substituents ortho to the hydroxy functions, as it is the case of tetra-halogenated BPA (TBBPA and TCBPA) and tetra-methyl BPA (TMBPA), are not expected to be prone to follow this ortho-oxidation pathway.

A second possibility is an oxidation taking place opposite to the hydroxy function, and involving a P450 mediated ipso-addition, immediately followed by an ipso-substitution rearrangement. For chemical reasons, this pathway can occur only for bisphenols that possess a central quaternary carbon.

It was clearly demonstrated for BPA (Jaeg et al. 2004), with its molecular mechanisms fully detailed by Nakamura et al. (2011). Ultimately, this peculiar pathway leads to a cleavage of the BPA molecule into two smaller metabolites, namely isopropenyl-phenol and hydroxycumyl alcohol, the latter having also been suggested to be an active metabolite at the level of endoplasmic reticulum (Nakamura et al. 2011). This first-step oxidation position is also concomitant with the production of glutathione conjugates hinting for the formation of reactive intermediates (Jaeg et al. 2004), as well as the formation of dimeric structures (Jaeg et al. 2004; Yoshihara et al. 2004). This metabolic pathway was also demonstrated to occur for TBBPA (Zalko et al. 2006). Notably, based on TBBPA studies, it is likely that this pathway is enhanced by the presence of halogens (Zalko et al. 2006). Data still lack about many BPA alternatives for which there are fair chances this pathway occurs, i.e. all molecules which possess a central quaternary carbon (e.g. BPBP, BPAP, and bisphenol BH (BPPH)), and particularly mono- to tetra-halogenated BPA, including TCBPA. The further glucuronidation or sulfation of these hydroxylated metabolites was also reported in vitro for BPA as well as other analogues such as BPS and BPF (Cabaton et al. 2008; Skledar et al. 2016; Gys et al. 2018).

Finally, it should be mentioned that bisphenols oxidation can also occur at the level of a methyl group next to the central carbon, as first demonstrated for BPA (Zalko et al. 2003). This may occur for any analogue in which the central carbon bears one or two methyl substituents (e.g. BPE, BPB, BPM, BPAP, and halogenated BPA analogues). Although data lack for alternatives other than BPA, it is unlikely that this specific oxidation position would result in a bioactivation process, conversely to the previously detailed pathways.

## 2.6. Isoform specific metabolism of BPA analogues

Besides the above-described major knowledge gaps that still exist for several BPA alternatives, additional information about isoform-specific metabolism (i.e. identification of the human isozymes involved in BPA alternatives metabolism) with the related biochemical parameters (including the intrinsic clearance) is also missing for most substances. This kind of data can be obtained by using human-derived in vitro models (e.g. recombinant enzymes, human microsomes, and other sub-cellular fractions from various organs) and gives a fundamental input for PBK modeling to estimate internal dose and identify human variability (Testai et al. 2021). As regards phase II metabolism, available data suggest that BPA analogues are detoxified by different UDP glucuronosyltransferase (UGT) isoforms with respect to BPA, that is UGT2B15, followed by UGT1A9, UGT2B7, UGT1A1, and UGT2B4 (Hanioka et al. 2008). The glucuronidation of BPAF, at 10 µM, was assessed using 12 human recombinant UGTs. UGT2B7 was found to be the most active, while no BPAF glucuronide was formed when using UGTs 1A4, 1A6, 1A7, and 1A10 (Li M et al. 2013). More recently testing of a range of BPAF concentrations (0.1–50 µM), a1A3, 2B7, 2B17, 1A10, and 2A1 showed similar intrinsic clearance value (calculated as the Vmax/Km
<figure><img src="image_21.png" /></figure>

ratio), although the ones with the highest affinity, i.e. 2B7 and 2B17 were found to be the most relevant at low exposure concentration (Skledar et al. 2019). Of note, UGT2B17, besides being a hepatic isoform, also accounts for >50% of the total intestinal UGTs (Kasteel et al. 2020), indicating that the pre-systemic detoxification, limiting the bioavailability of the toxic parent, can be extremely relevant. Other isoforms were mainly responsible for BPS (UGT1A9) and BPF glucuronidation (UGT1A10) (Skledar et al. 2015). No data are available on the isoform specific glucuronidation of the BPA alternatives selected in the PARC project. However, since the available data seem to indicate differences among similar molecules, it is worthwhile to investigate it, considering the interindividual differences as well as differences in the life stages, since some UGT isoforms are not fully expressed in the neonatal period.

The metabolism of BPS, BPF, and BPAF was also investigated using human liver microsomes (HLM) and human intestine microsomes (HIM), showing that the intestine contributes less than the liver to their overall metabolism (Skledar et al. 2016). It was also demonstrated that the relative efficiency of detoxification was highly dependent upon the exposure concentration: at lower substrate concentrations, representative of human exposure, the glucuronidation of BPAF was the most effective while that of BPS was the least effective among all the tested analogues (Karrer et al. 2018).

Sulfation of BPA is catalyzed predominantly by the sulfo-transferase (SULT) SULT1A1, SULT2A1, and SULT1E1 (Suiko et al. 2000; Nishiyama et al. 2002). For BPA analogues, no data on single recombinant SULTs are available for any bisphenol to explore the isoform specificity of this reaction. However, sulfation was investigated for BPA, BPS, and BPF using human liver cytosol (HLC). By comparing sulfation with previously reported kinetic parameters for glucuronidation, it was shown that BPA and BPF sulfation was less efficient than glucuronidation, while for BPS the intrinsic clearance was closer between the two reactions (Durcik, Skledar, et al. 2022).

For phase I oxidative pathways, only limited data are available, and studies have focused only on the CYP-mediated formation of hydroxylated metabolites. In addition to the study of Nakamura et al. (2011), highlighting the role of the CYP3A and CYP2D6 isoforms, Schmidt et al. (2013) compared the formation of hydroxylated metabolites for BPA and BPF, BPAF, and BPZ using single CYPs isoforms (CYP3A4, CYP1A2, CYP2C8, CYP2C9, CYP2C19, and CYP2E1) and a single substrate concentration. All CYPs tested were active, suggesting that hydroxylation reactions are not specifically catalyzed by one specific CYP and ipso-substitution reactions of bisphenols were not found to be major pathways in this study. However, the production of reactive intermediates through this specific pathway may lead to their immediate binding to endogenous macromolecules and/or to the cleavage of the parent molecule into smaller metabolites, therefore, complicating the qualitative and quantitative monitoring of the contribution of the ipso-substitution pathway in bisphenols metabolism. Despite bisphenols structural proximity, differences were found to exist: CYP1A2 and CYP2C9 mainly catalyze

BPA and BPF hydroxylation, whereas BPAF is mainly hydroxylated by CYP3A4 and to a lower extent by CYP2C9, BPZ is mainly metabolized by CYP2C8 (Schmidt et al. 2013); BPS has been reported to be hydroxylated by CYP3A4 and CYP2C9 (Skledar et al. 2016). However, neither kinetic parameters nor their dependence on the concentration of the different bisphenols were derived.

## 2.7. Halogenated bisphenols and other specific alternatives

Halogenated bisphenols are a key sub-family of analogues, which mainly comprises TBBPA, a substance that still accounts for 50% of the flame retardants market, but also chlorinated and mixed Br/Cl bisphenols. The kinetics of TBBPA in humans have been reported following a single oral dose of 0.1 mg/kg bw TBBPA (five individuals) and in rats administered a single oral dose of 300 mg/kg bw TBBPA (Schauer et al. 2006). TBBPA was not present at detectable concentrations in any of the human plasma samples, while TBBPA-glucuronide and TBBPA-sulfate were identified as metabolites of TBBPA in blood and urine of both humans and rats, suggesting a low systemic bioavailability of the parent compound. In humans, a higher frequency of detection and levels were reported for the TBBPA-glucuronide. The absorption, as well as the conjugation, were rapid processes ($C_{max}$ for the glucuronide was <4 h in humans and <3 h in rats), whereas the urinary excretion was slow. In rats, in addition to TBBPA-glucuronide and sulfate, a diglucuronide of TBBPA, a mixed glucuronide-sulfate conjugate of TBBPA, tri-bromo bisphenol A, and the glucuronide of tri-bromo bisphenol A were also present in low concentrations. TBBPA in rat plasma showed a half-life of 13 h (Schauer et al. 2006). Low concentrations of TBBPA have been shown to inhibit BCRP-mediated transport in rat brain capillaries (Cannon et al. 2019). This information is relevant since blocking BCRP-mediated efflux is likely to result in higher chemical exposure of the brain.

Halogenation likely enhances the potential of such bisphenols to undergo metabolic activation through phase I reactions, with ROS production and possible links with induced inflammation, immunotoxicity, and genotoxicity. In addition, TCBPA, TBBPA as well as TBBPA mono-sulfate were demonstrated to be potent activators of the peroxisome proliferator-activated receptor gamma (PPARγ) (Riu, Le Maire, et al. 2011) unlike most other BPA analogues, which primarily target an ER response. Thus, halogenated BPA may trigger a modulation of energy metabolism through a distinct mode of action.

Alternatives used in thermal papers (structurally close to bisphenols: BPS-MAE and BPS-isopropyl (BPS-IP), or less close: Pergafast-201) form another sub-family of concern. Little is known about their fate, especially as regards the Pergafast family. Note that compounds like Pergafast-201 are also expected to break down into two daughter molecules in solution, therefore, requiring the study of by-products in addition to the mother molecule (Eckardt et al. 2020).
<figure><img src="image_22.png" /><figcaption>**Table 2.** Synthesis of the range or single IC~50~/AC~20~ values reported in the literature for each BPA alternative, grouped by *in vitro* assay categories.</figcaption></figure>

<table><thead><tr><th></th><th>BPAP</th><th>BPE</th><th>BPP</th><th>BPS-MAE</th><th>BPZ</th><th>TBBPA</th><th>TCBPA</th></tr></thead><tbody><tr><td>Erα binding affinity</td><td>255 nM to 2.6 μM</td><td>5.3 μM</td><td>152 nM to 0.671 μM</td><td>Inactive</td><td>11–80 nM</td><td>250 nM</td><td>2.8 μM</td></tr><tr><td>Erα reporter gene assay agonism</td><td>351 nM to 2 μM</td><td>0.102–7.43 μM</td><td>0.641 μM (RIC<sub>20</sub>)</td><td>Inactive</td><td>3.7 pM to 8.63 μM</td><td>Inactive/19 μM</td><td>20 nM to 47.60 μM</td></tr><tr><td>Antagonism</td><td>24.54 nM (IC<sub>20</sub>)</td><td></td><td>3.8 nM (RIC<sub>20</sub>) to 1.91 μM</td><td></td><td></td><td>0.6 μM</td><td></td></tr><tr><td>E-screen (proliferation)</td><td>0.39 μM</td><td>0.1–10 μM</td><td>NA</td><td>NA</td><td>0.11 μM</td><td>124 μM</td><td>45.8 μM</td></tr><tr><td>Erα gene expression</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>↑ 1 pM</td><td>↑ 1–10 μM/↓ 25 μM</td></tr><tr><td>Erα protein expression</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>Inactive</td><td>↑ 10 nM to 1 μM</td></tr><tr><td>Erβ binding affinity</td><td>62.1–68 nM</td><td>3.9 μM</td><td>105 nM</td><td>Inactive</td><td>128 nM to 1.1 μM</td><td>Inactive</td><td>NA</td></tr><tr><td>Erβ reporter gene assay agonism</td><td>0.17–0.61 μM</td><td>94 nM to 4.6 μM</td><td></td><td>Inactive</td><td>18 nM to 0.5 μM</td><td></td><td>1.15–68 μM</td></tr><tr><td>Antagonism</td><td>595 nM</td><td></td><td>0.93–1.81 μM</td><td></td><td>1.16 μM</td><td>↑ 100 pM to 10 μM</td><td></td></tr><tr><td>ERβ gene expression</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td></tr><tr><td>ERRγ binding affinity</td><td>120 nM</td><td>4.78 nM</td><td>320 nM</td><td>NA</td><td>13.4 nM</td><td>Inactive</td><td>Inactive</td></tr><tr><td>ERRγ reporter gene assay</td><td>Inactive</td><td>0.69 μM</td><td>Inactive</td><td>NA</td><td>5.24 μM</td><td>↑ 100 pM</td><td>NA</td></tr><tr><td>ERRγ gene expression</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>↑ 10–50 nM</td><td>Inactive</td></tr><tr><td>GPCR proliferation assay</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>↑ 1 pM to 10 μM</td><td>↑ 10 nM to 10 μM</td></tr><tr><td>GPCR gene expression</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>↑ 10 nM to 10 μM</td><td>↓ 1 nM/↑ 10 nM to 1 μM</td></tr><tr><td>GPCR protein expression</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>↑ 10 nM to 10 μM</td><td>NA</td></tr><tr><td>AR binding affinity</td><td>5.4–>10 μM</td><td>>10 μM</td><td>409 nM</td><td>Inactive</td><td>1.6–6.68</td><td>1–10 μM</td><td>100 nM (RIC<sub>20</sub>) to 13.4 μM</td></tr><tr><td>AR reporter gene assay antagonism</td><td>8.5–40 μM</td><td>0.99–10 μM</td><td>2.0–10.8 μM</td><td>Inactive</td><td>3.58 nM to 7.3 μM</td><td>0.4–0.982 μM</td><td>Inactive</td></tr><tr><td>AR anti-proliferative assay</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td></tr><tr><td>AR protein expression</td><td>↓ 16.0 μM</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td></tr><tr><td>PR binding affinity</td><td>330 nM</td><td>inactive</td><td>896 nM</td><td>NA</td><td>1.89 μM</td><td>NA</td><td>NA</td></tr><tr><td>PR reporter gene assay antagonism</td><td>23.1 μM</td><td>inactive</td><td>7.18 μM</td><td>NA</td><td>7.64 μM</td><td>NA</td><td>78 nM (RIC<sub>20</sub>)</td></tr><tr><td>PR gene expression</td><td>↑ 1 μM</td><td>↑ 1 μM</td><td>↑ 1 μM</td><td>NA</td><td>↑ 1 μM</td><td>↑ 30 μM</td><td>↑ 1–25 μM/↓ 10 nM to 10 μM</td></tr><tr><td>PR protein expression</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>0.5 μM</td><td>NA</td></tr><tr><td>GR binding affinity</td><td>30.1 nM</td><td>>10 μM</td><td>213 nM</td><td>NA</td><td>226 nM</td><td>Inactive</td><td>Inactive</td></tr><tr><td>GR reporter gene assay agonism</td><td>158 nM</td><td>Inactive/↑</td><td>0.32 nM (LOEC)</td><td>NA</td><td>1.99 μM (LOEC)</td><td>22 nM</td><td>NA</td></tr><tr><td>Antagonism</td><td>8.1 μM</td><td></td><td>17.8 nM (RIC<sub>20</sub>) to 16.1 μM</td><td></td><td>8.8–22 μM</td><td>↑ 1–100 pM</td><td>NA</td></tr><tr><td>GR gene expression</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>0.7–5.7 μM</td><td>6 μM</td></tr><tr><td>PPARγ binding affinity</td><td>>10 μM</td><td>Inactive</td><td>Inactive</td><td>NA</td><td>Inactive</td><td>0.2–397 μM</td><td>0.3–1.28 μM</td></tr><tr><td>PPARγ reporter gene assay agonism</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>↑ 1–20 μM</td><td>NA</td></tr><tr><td>PPARγ gene expression</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>↓ 10 μM/↑ 10 nM to 30 μM</td><td>↑ 1 nM</td></tr><tr><td>PPARα/β/δ binding affinity</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>68 μM</td><td>102 μM</td></tr><tr><td>PPARα/β/δ reporter gene assay</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>25–50 μM</td><td>50 μM</td></tr><tr><td>PPARα gene expression</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>↑ 1 μM</td><td>NA</td></tr><tr><td>PPARβ/δ gene expression</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td><td>↑ 1 μM</td><td>NA</td></tr><tr><td>PXR binding affinity</td><td>1.06 μM</td><td>>10 μM</td><td>2.45 μM</td><td>NA</td><td>2.15 μM</td><td>NA</td><td>NA</td></tr><tr><td>PXR reporter gene assay agonism</td><td>9.7 (RIC<sub>20</sub>) to 65.9 μM</td><td>Inactive</td><td>6.3 (RIC<sub>20</sub>) to 19 μM</td><td>NA</td><td>8.7–10.2 μM</td><td>11.97–123.1 μM</td><td>8.49–24.1 μM</td></tr></tbody></table>

AR: androgen receptor; ER: estrogen-related receptor; GPER: membrane G protein-coupled ER; GR: glucocorticoid receptor; IC: inhibitory concentration; LOEC: lowest observed effective concentration; NA: not assessed; PXR: pregnane X receptor; PPAR: peroxisome proliferator activated receptors; PR: progesterone receptor; REC: relative effective concentration; RIC: relative inhibitory concentration.

In case of IC~20~/AC~20~ values, they are specified; in addition, values related to genes/protein expression are concentrations at which significant effects were observed. The detailed list of all the assays and the related publications is shown in Table A1.
<figure><img src="image_23.png" /></figure>

## 2.8. Concluding considerations on the kinetics of BPA alternatives

Apart from the bioactivation issues that remain unexplored for many bisphenol alternatives, a better knowledge of the metabolic pathways of these substances may help in (i) decreasing the uncertainties in risk assessment of BPA alternatives, accounting for possible species differences between human and data obtained with animal models, following the example of BPA; (ii) providing key information to modelers, about the processes likely to impact actual internal biologically active doses; and (iii) identifying new candidate biomarker metabolites that have the potential to be used as exposure biomarkers to enhance human biomonitoring strategies. Of note, exposure to bisphenol alternatives may also result in a modulation of liver metabolic capacities, by inducing or repressing phase I and II xenobiotic metabolizing enzymes (XME) expression and functionality. This may well be a mode of action for these candidate metabolic disruptors, since these enzymes are also deeply involved in endogenous metabolisms, including that of many hormones. In that sense, bisphenols and their metabolites may act as metabolism disrupting chemicals (MDC) not only through the activation of specific receptors (ER, PPAR), but also through modulating liver XME activities.

In vitro systems are increasingly used to characterize the toxicological profile of chemicals, including bisphenol alternatives, and to collect information on their mechanism of action. However, frequently the metabolic competence and/or the activities of transporters in these systems are unknown. In vitro assays are generally carried out without considering the in vitro kinetics of the tested substance, which would enable the measurement of the actual intracellular concentration of the active/toxic compound(s). Consequently, the nominal concentration is typically used for in vitro dose-response curves (Kramer et al. 2015). How the use of the nominal concentrations could lead to the misinterpretation of results has been reported in a number of publications (Wilmes et al. 2013; Kramer et al. 2015; Pomponio et al. 2015; Truisi et al. 2015; Gouliarmou et al. 2018); in addition, the OECD, among others, has clearly pointed out that the metabolic characterization of in vitro systems and biokinetics measurements are pivotal requirements, in its Guidance Document on Good In Vitro Method Practices (GIVIMP) (OECD 2018). The importance of biotransformation processes in determining the bioavailable concentration of parent bisphenols is highlighted by the fact that only the parent substances (in most cases), but not their conjugates, are able to activate nuclear receptors (NRs). These considerations as well as the potential formation of reactive metabolites, should be considered any time an in vitro system is used when interpreting data and trying to extrapolate to the actual in vivo situation. Therefore, the observed in vitro effects should be interpreted with caution, if the intracellular context of exposure (actual concentration, presence of metabolites) was not assessed, especially in the absence of effects.

# 3. Endocrine disrupting activity of BPA alternatives

## 3.1. General considerations

Based on the WHO definition of EDCs (WHO 2013), the European Commission established three criteria that a substance should fulfill for ED classification. These criteria are related to endocrine activity, adverse effect and the biological plausible link between the endocrine activity and the adverse effect (ECHA and EFSA 2018). BPA has been classified as SVHC for its ED activity. Despite data have been already produced and analyzed for ED-related adverse effects of several BPA alternatives (Pelch KE et al. 2017; Pelch K et al. 2019), the following section aims to update and describe the currently available evidence on BPA alternatives' ED activity on NRs, thyroid hormones (THs) homeostasis and steroidogenesis. Only publications using chemical standards of compounds identified as priority in PARC (i.e. BPE, BPP, BPZ, BPAP, BPS-MAE, and TCBPA) have been considered, excluding assessment of ED activity of matrix extracts of any origin (thermal paper, sludge, etc.). Studies describing effects of TBBPA were also considered for comparison due to chemical similarity with TCBPA.

## 3.2. Interaction and effects on nuclear steroid, PPARs, and other receptors

The following sections describe effects exerted by BPA alternatives on these NRs based on in vitro studies. Data are summarized in Table 2 and reported in detail in Table A1.

### 3.2.1. Estrogen receptors (ERα, ERβ, ERRγ, GPER)

The estrogenic activity of BPA is certainly the best characterized; therefore, effects of BPA alternatives on nuclear estrogen receptors (ERα, ERβ) are the most investigated. The relative binding affinities of TCBPA and TBBPA, assessed in competitive binding experiments on ERα isolated from breast MCF-7 cells, were much lower than that of BPA (Samuelsen et al. 2001; Olsen et al. 2003). By using ERα affinity columns, only TCBPA bound to the receptor (Riu, Le Maire, et al. 2011).

In a gene reporter assay in MCF-7 cells, TCBPA displayed stronger estrogenic activity than BPA while TBBPA had weak activity. In contrast, TBBPA 10 µM exhibited anti-estrogenic activity in presence of E2 10 pM (Kitamura et al. 2005) and in a yeast two-hybrid assay (van Leeuwen et al. 2019). However, TBBPA did not display any agonist/antagonist activity in several other studies (Hamers et al. 2006; Li J et al. 2010; Lee HK et al. 2012; Molina-Molina et al. 2013; Ruan et al. 2015; Grimaldi et al. 2019; Durcik, Hiti, et al. 2022). TCBPA had no activity in a two-hybrid yeast assay (Li J et al. 2010), whereas it was confirmed to be an ERα agonist using stably transfected cell lines (Riu, Grimaldi, et al. 2011; Riu, Le Maire, et al. 2011; Molina-Molina et al. 2013; Pelch KE et al. 2019; Durcik, Hiti, et al. 2022), and in a bioluminescence yeast estrogen screen assay (Ruan et al. 2015) showing a potency generally lower than that of BPA. TCBPA had agonist potency lower than BPA also for ERβ in transfected HGELN-ERβ cells (Riu, Le Maire, et al. 2011) while TBBPA was inactive (Pelch KE et al.
2019) or had very limited effects on both ERs (Riu, Grimaldi, et al. 2011; Riu, Le Maire, et al. 2011).

BPZ displayed a dose-dependent binding to ERα comparable to BPA (Böckers et al. 2020). BPAP and BPZ, showed to be highly selective, with BPAP having a higher affinity for ERβ and BPZ for ERα (Liu X et al. 2019, 2021). Another report confirmed the higher affinity of BPZ for ERα while no binding to ERβ was observed for BPAP (Keminer et al. 2020). BPZ had a stronger binding affinity than BPP for ERα in MVLN cells, whereas BPAP was inactive (Lin et al. 2021). BPE was a weak binder of ERα in a yeast two-hybrid assay, while displaying a strong binding activity in a cell-free system (Hashimoto et al. 2001). In addition, BPE and BPP had high binding affinity for both ERs, with BPE being more potent; interestingly, BPE was also a strong binder for ERRγ, with a potency similar to that of BPA. The affinity of the other alternatives for ERRγ was lower, with the following decreasing order: BPZ > BPAP > BPP (Liu X et al. 2019).

Agonistic effects of BPE for ERα were observed in several *in vitro* models, with a potency often comparable or higher than BPA (Rosenmai et al. 2014; Ruan et al. 2015; Conroy-Ben et al. 2018; van Leeuwen et al. 2019; Chen Q et al. 2020; Durcik, Hiti, et al. 2022). Similar results were observed also in transfected cells for both ERs (Grimaldi et al. 2019; Kojima et al. 2019; Pelch KE et al. 2019).

BPZ had similar or higher agonist potency as BPA for ERα in transfected MCF-7 cells and yeast assays (Mesnage et al. 2017; Conroy-Ben et al. 2018; Pelch KE et al. 2019; van Leeuwen et al. 2019). Conversely, BPZ was less potent than BPA for ERβ in transfected HepG2 cells (Pelch KE et al. 2019), whereas in other transfected cells BPZ was a strong agonist for both ERs (Grimaldi et al. 2019; Kojima et al. 2019). The transactivation potency for ERα was confirmed in other reporter gene assays but BPZ was almost inactive for ERβ (Chen Q et al. 2020; Keminer et al. 2020; Liu X et al. 2021) or, rather with antagonist activity (Liu X et al. 2021). Opposite evidence was observed in human embryonic kidney HEK293 cells over-expressing ERα, where BPZ (25 µM) reduced ERα activity with time (Böckers et al. 2020).

In various reporter gene assays, BPAP had a lower agonist potency than BPA for ERα (Mesnage et al. 2017; Grimaldi et al. 2019; Kojima et al. 2019; Pelch KE et al. 2019; Liu X et al. 2021), whereas it was completely inactive in a yeast two-hybrid assay (van Leeuwen et al. 2019). However, BPAP displayed antagonist activity for ERα in MVLN cells (Chen Q et al. 2020), as well as agonist (Grimaldi et al. 2019; Kojima et al. 2019) or antagonist activity (Liu X et al. 2021) for ERβ using different transfected cells.

BPP was inactive for ERα in two different *in vitro* systems (Pelch KE et al. 2019; van Leeuwen et al. 2019) whereas it exhibited antagonist activity for ERα in reporter gene assays (Grimaldi et al. 2019; Chen Q et al. 2020). BPP was also antagonist of ERβ in transfected cells (Grimaldi et al. 2019; Pelch KE et al. 2019). BPP agonism for ERα was observed only up to 1 µM, then decreasing its activity with the dose, but it was a strong antagonist for both ERs (Kojima et al. 2019).

BPS-MAE has been limitedly investigated, with data reporting its inactivity as binder or as agonist/antagonist on both receptors (Pelch KE et al. 2019; Keminer et al. 2020).

BPA has a high affinity for ERRγ (Takayanagi et al. 2006). In transfected HELN-ERRγ cells, BPE had the highest potency, followed by BPZ, whereas BPAP, BPP, TCBPA, and TBBPA were inactive (Grimaldi et al. 2019), confirming for the latter two compounds what was previously observed in a yeast assay (Li J et al. 2010).

By the E-screen assay, assessing the proliferation induction of MCF-7 ERα positive cells, TCBPA displayed a very low proliferative potential, only at the highest concentration tested, while TBBPA had no effects (Samuelsen et al. 2001; Olsen et al. 2003; Molina-Molina et al. 2013). However, in another report, TBBPA exerted both estrogenic and anti-estrogenic potential (in presence of E2) on MCF-7 cell growth (Krivoshiev et al. 2016). In contrast, BPE, BPZ, and BPAP strongly promoted cell growth (Hashimoto et al. 2001; Mesnage et al. 2017). To our knowledge, no evidence is available on the proliferative potential of BPP and BPS-MAE. Neither TCBPA (10 µM) nor TBBPA (30 µM) affected ERα protein expression in MCF-7 cells (Olsen et al. 2003). However, in same cells, TCBPA induced ERα protein expression in the range 10 nM to 1 µM, whereas gene expression was induced at 1 and 10 µM and repressed at 25 µM (Lei, Tang, et al. 2021). In mouse 3T3-L1 adipocytes, TBBPA induced the gene expression of ERα only at 1 pM, of ERβ up to 10 µM and of ERRγ at 100 pM (Chappell et al. 2018). No evidence on modulation of ERα (except TCBPA), ERβ or ERRγ gene/protein expression by TCBPA, BPE, BPP, BPZ, BPAP, and BPS-MAE is available.

Effects on the membrane G protein-coupled ER (GPER, or GPR30), mediating non-genomic estrogen signaling and recognized as a BPA target (Cimmino et al. 2019) were assessed only for the halogenated BPA alternatives. In particular, proliferation of ovarian OVCAR-3 and granulosa KGN cells was promoted by TBBPA through the GPER pathway, whereas TCBPA did not affect cell growth (Hoffmann et al. 2017). In same cells, both compounds did not affect GPER gene expression (Hoffmann et al. 2017). In mouse MA-10 Leydig cells, both BPA and TCBPA (1 nM) repressed GPER protein expression, whereas TBBPA (10 nM) significantly induced it (Gorowska-Wojtowicz et al. 2019). However, in breast cell lines, TCBPA induced GPER gene and protein expression (Lei, Tang, et al. 2021; Yu M et al. 2023).

In summary, despite the difference in the methods and cellular models used, the collected evidence highlights that the two halogenated compounds are very different in their estrogenic activity, with TBBPA being almost inactive and TCBPA having agonist activity toward both ERs. The non-halogenated BPA alternatives displayed a spectrum of different activities: BPE and BPZ had the highest agonist potency toward all the ERs, including ERRγ, whereas BPAP had a moderate activity; BPAP seemed more active for ERβ and BPZ for ERα; BPP had a stronger antagonist activity for both ERs and no activity for ERRγ. BPS-MAE appeared to be inactive but more data are needed to substantiate this evidence.

### 3.2.2. Androgen receptor (AR)

BPA is known to antagonize AR (Rubin 2011). TBBPA dose-dependently displaced testosterone from AR (Beck et al.
<figure><img src="image_24.png" /></figure>

2016). BPP and BPZ had a moderate and low activity, respect-
ively, whereas BPA and BPE were not active (Liu X et al.
2019). No data on binding of TCBPA, BPAP, and BPS-MAE are
available.

Different transactivation assays were performed to evalu-
ate agonist/antagonist activity of BPA alternatives and none
of them evidenced agonistic activity for AR. Contrasting evi-
dence is available for halogenated derivatives as regards their
anti-androgenic activity. In some cells stably transfected with
AR, no activity was detected for TCBPA and TBBPA (Kitamura
et al. 2005; Molina-Molina et al. 2013). TBBPA was not active
also in AR-CALUX and two-hybrid yeast assays (Hamers et al.
2006; Li J et al. 2010). Opposite evidence was reported for
TCBPA displaying anti-androgenic activity in a gene reporter
assay on kidney fibroblast CV-1 cells in presence of 5α-
dihydrotestosterone (DHT) 1 nM (Sun et al. 2006), or in a two-
hybrid AR yeast assay (Li J et al. 2010). TBBPA antagonism
was observed with a yeast receptor bioassay (Roelofs et al.
2015), and in transfected HEK-293 cells with TBBPA (10 μM)
reducing AR activation by 37% in presence of testosterone
(Beck et al. 2016).

In transfected HELN cells, BPA, BPE, BPZ, BPP, TCBPA, and BPAP, but not TBBPA, displayed AR antagonism, with different potencies (Grimaldi et al. 2019). BPE had a stronger antagonistic effect than BPA in various reporter or yeast assays (Rosenmai et al. 2014; Conroy-Ben et al. 2018; Šauer et al. 2021). In transfected Chinese hamster ovary (CHO)-1 cells, the following anti-androgenic potency was found: BPE > BPA > BPZ, BPP, and BPAP (Kojima et al. 2019); BPZ failed to show antagonism in a yeast assay (Conroy-Ben et al. 2018). Conversely, in transfected HepG2 cells, BPA, BPZ, and BPE had antagonistic effects with comparable high potencies; BPP was less potent, whereas BPAP and BPS-MAE did not show any significant activity (Pelch KE et al. 2019). BPE, BPZ, BPP, and BPAP showed anti-androgenic effects also in transfected MDA-kb2 mammary cells (Kolšek et al. 2015; Ma M et al. 2022), whereas in an AR-CALUX assay BPP did not show significant anti-androgenic activity (Šauer et al. 2021). BPS-MAE had no effect (Pelch KE et al. 2019).

TCBPA and TBBPA did not affect proliferation of MCF-7
cells transfected with AR (Molina-Molina et al. 2013). No evi-
dence is available for BPE, BPP, BPZ, BPAP, and BPS-MAE.
BPAP at 10 µM decreased AR protein nuclear expression in
22Rv1, a prostate cell line, as well as in MCF-7/LLC2 cells in
presence of DHT, but not in prostate LNCaP cells. Similarly,
BPAP decreased AR gene expression and cell growth in
22Rv1 cells (Stossi et al. 2016).

Overall, TCBPA and TBBPA are not strong AR antagonists,
and the following potency can be summarized for the other
alternatives BPE > BPZ > BPP > BPAP; BPS-MAE is inactive,
but this finding is supported by only one study.

**3.2.3. Progesterone receptor (PR)**

BPP and BPAP showed moderate binding activity toward PR,
BPZ was very weak and BPE completely inactive, like BPA (Liu
X et al. 2019). No report on TCBPA, TBBPA, and BPA-MAE
binding affinity is available.

No agonist activity of BPA and its alternatives toward PR
was evidenced in transfected HELN-PR cells. Rather, BPP, BPZ,
and BPAP displayed strong antagonistic activity while BPE,
TCBPA, and TBBPA were inactive as BPA (Grimaldi et al.
2019). In PR-CALUX assays, TBBPA, BPE, and BPP did not
show antagonist activity (Hamers et al. 2006; Šauer et al.
2021). Conversely, TCBPA and TBBPA displayed some antag-
onist activity in a two-hybrid yeast assay (Li J et al. 2010).

In MCF-7 cells, TCBPA induced PR gene expression in the range 10 nM to 1 µM (Lei, Tang, et al. 2021) or protein expression at 10 µM, to an extent comparable to BPA 300 µM, while TBBPA 30 µM had a lower effect (Samuelsen et al. 2001; Olsen et al. 2003). On the contrary, in breast SKBR3 cells, TCBPA repressed PR gene expression in the range 10 nM to 10 µM (Yu M et al. 2023). In another study on MCF-7 cells, TCBPA did not affect PR gene expression at 1 µM; otherwise, at same concentration, BPA, BPZ, BPE, BPP, and BPAP all induced PR gene expression (Pelch KE et al. 2019).

Overall, no definitive conclusion could be drawn for
TCBPA, TBBPA, BPP, BPZ, and BPA on their interaction with
PR due to the contrasting evidence. BPE is the only com-
pound appearing quite inactive. No evidence is available for
BPS-MAE.

**3.2.4. Glucocorticoid receptor (GR)**

In two different competitive binding assays, TBBPA had no affinity (Beck et al. 2016) or strong affinity for GR (Liu QS et al. 2020); BPP and BPZ had a strong binding affinity for GR, fivefold higher than BPA, whereas BPAP was moderately active (Liu X et al. 2019). No data on binding affinity is available for TCBPA, BPE, and BPS-MAE.

In transfected HMLN-GR cells, BPP was a potent GR antagon-
onist while BPA, BPZ, BPAP, BPE, TCBPA, and TBBPA had no
activity (Grimaldi et al. 2019). Similarly, TBBPA was inactive in
two transfected models (Beck et al. 2016; Chappell et al.
2018). However, by using a yeast assay, TBBPA displayed a
strong GR antagonism (Roelofs et al. 2015) and in transfected
CHO-K1 cells, BPA, BPAP, BPP, and BPZ were all antagonists
and BPE inactive (Kojima et al. 2019). BPAP may exert GR
agonist activity in MDA-kb2 cells, whereas BPA, BPP, and BPZ
were agonist in MDA-kb2 and CHO-K1 cells; in addition, BPA
was antagonist in both cells whereas BPP and BPZ were
antagonist in CHO-K1 and MDA-kb2 cells, respectively. BPE
confirmed to have no activity on GR (Chen Q et al. 2020).
However, in studies using transfected MDA-kb2 cells, BPE
exerted GR agonism while BPA, BPP, BPAP, and BPZ were
antagonists (Kolšek et al. 2015; Ma M et al. 2022).

In one report using 3T3-L1 cells, TBBPA induced GR gene
expression (Chappell et al. 2018). No evidence is available for
the other BPA alternatives.

In summary, TCBPA seemed to be inactive but was investi-
gated in only one report. Also, TBBPA and BPE appeared to
have no clear activity toward GR. Evidence supports BPP and
BPZ as GR antagonists whereas contrasting findings are avail-
able for BPAP. No data are available for BPS-MAE.
<figure><img src="image_25.png" /><figcaption>**Table 3.** Summary of the range of active concentrations reported in the literature for each BPA alternative, grouped by *in vitro* assay categories as regard thyroid-disrupting properties.</figcaption></figure>

<table><thead><tr><th></th><th>BPAP</th><th>BPE</th><th>BPP</th><th>BPS-MAE</th><th>BPZ</th><th>TBBPA</th><th>TCBPA</th></tr></thead><tbody><tr><td>TRα binding affinity</td><td></td><td></td><td></td><td></td><td></td><td>IC<sub>20</sub> 1.5 µM (RBP/<br/>T3 = 0.07)</td><td></td></tr><tr><td>TRα gene reporter assay</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Agonism</td><td></td><td>Potency 70 µM</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Antagonism</td><td></td><td></td><td></td><td></td><td></td><td>10–60 µM<br/>10–100 µM</td><td>NA/weak agonist</td></tr><tr><td>GH3 cell proliferation</td><td>✓1 µM</td><td></td><td>✓1 µM</td><td></td><td>✓1 µM</td><td>✓1 µM<br/>✓100 µM</td><td>✓100 µM</td></tr><tr><td>THRβ binding affinity</td><td></td><td></td><td></td><td></td><td></td><td>IC<sub>20</sub> 5 µM (RBP/<br/>T3 = 0.01)</td><td></td></tr><tr><td>THRβ gene reporter assay</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Agonism</td><td></td><td>NA</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Antagonism</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Expression of TH-target genes in vitro (GH3 cells)</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Tshβ</td><td>NA</td><td></td><td>NA</td><td></td><td>NA</td><td></td><td></td></tr><tr><td>Thrα</td><td>✓(34 µM)</td><td></td><td>NA</td><td></td><td>✓(0.37 µM)</td><td></td><td></td></tr><tr><td>Thrβ</td><td>✓(34 µM)</td><td></td><td>✓(29 µM)</td><td></td><td>✓(3.7 µM)</td><td></td><td></td></tr><tr><td>Dio1</td><td>NA</td><td></td><td>✓(29 µM)</td><td></td><td>NA</td><td></td><td></td></tr><tr><td>Dio2</td><td>✓(34 µM)</td><td></td><td>✓(29 µM)</td><td></td><td>✓(37 µM)</td><td></td><td></td></tr><tr><td>TTR binding</td><td></td><td>✓T4 binding IC<sub>50</sub><br/>0.13 mM</td><td></td><td></td><td></td><td>✓T4 binding IC<sub>50</sub><br/>7.7 nM (RBP/T4<br/>1.5–10.6)</td><td>✓T4 binding IC<sub>50</sub><br/>107 nM (RBP/<br/>T4 0.76)</td></tr><tr><td>TBG binding</td><td></td><td></td><td></td><td></td><td></td><td>NA</td><td>NA</td></tr></tbody></table>

NA: not active.

Values related to genes/protein expression are concentrations at which significant effects were observed. The detailed list of all the assays and the related publications is shown in Table A2. When given in the article, relative binding potencies (RBP) toward the natural ligand (T3 for THR, T4 for TTR) are reported. Effects on the expression on TH synthesis-related genes are not reported as they rely on a single study (Lee S et al. 2017) and did show an effect of the tested compounds (BAP, BPP, and BPZ).

### 3.2.5. Peroxisome proliferator activated receptors (PPARα, β/δ, γ)

By using a whole-cell competitive assay or affinity columns, both TCBPA and TBBPA, but not BPA, showed to be strong binders of PPARγ (Riu, Grimaldi, et al. 2011; Riu, Le Maire, et al. 2011). TBBPA confirmed to have a high affinity for PPARγ (Liu QS et al. 2020), whereas BPA, BPE, BPP, BPZ, and BPAP were inert or inactive (Liu X et al. 2019). Both TBBPA and TCBPA exhibited about 18- and 12-fold stronger binding affinity than BPA to PPARβ/δ, respectively (Li CH et al. 2021). The other BPA alternatives were not investigated for their binding to PPARβ/δ and no compound was evaluated for PPARα affinity.

TCBPA and TBBPA exhibited a high potency in activating PPARγ gene expression in transfected cells (Riu, Grimaldi, et al. 2011; Garoche et al. 2021). In addition, TBBPA had agonist activity in cells transfected with mouse or human PPARγ (Akiyama et al. 2015; Watt and Schlezinger 2015; Chappell et al. 2018; Andrews et al. 2020; Liu QS et al. 2020). TCBPA and TBBPA showed agonistic activity also for PPARβ/δ (Li CH et al. 2021), whereas they did not activate PPARα and PPARβ using other reporter assays (Riu, Grimaldi, et al. 2011; Riu, Le Maire, et al. 2011). The other alternatives were not investigated for their capability to trans-activate PPARs, most probably for their lack of binding affinity, at least for PPARγ.

TBBPA induced PPARγ gene and/or protein expression in several *in vitro* models (Akiyama et al. 2015; Honkisz and Wójtowicz 2015b; Watt and Schlezinger 2015; Chappell et al. 2018; Kakutani et al. 2018; Liu QS et al. 2020; Kim S et al.

2021; Cheng and Volz 2022). In MA-10 mouse Leydig cells, TBBPA (10 nM) significantly induced PPARγ protein expression unaffecting PPARβ, whereas TCBPA at 1 nM significantly decreased the protein expression of PPARβ and PPARγ. Both increased the expression of PPARα, TCBPA to a higher extent (Gorowska-Wojtowicz et al. 2019). TBBPA also up-regulated PPARα gene expression in 3T3-L1 mouse adipocytes (Kim S et al. 2021) and PPARβ/δ transcript levels in steatotic FaO rat hepatoma cells while decreasing the expression of PPARγ and α (Grasselli et al. 2014) in these cells. A drop in PPARγ protein expression was also induced by TBBPA in mouse neocortical neurons (Wojtowicz et al. 2014), whereas it did not affect PPARγ gene expression in human bone marrow cells (Andrews et al. 2020). Overall, the available evidence points to TBBPA as the only BPA alternative exerting a strong PPARγ agonism, with cell type-dependent effect on PPARγ expression.

### 3.2.6. Other receptors

Some studies were performed to explore interaction of BPA alternatives with receptors involved in the metabolism of xenobiotics. BPAP, BPP, and BPZ had a strong binding affinity for the pregnane X receptor (PXR), higher than that of BPA, whereas BPE was inactive (Liu X et al. 2019). In transfected HepG2 cells, TCBPA and TBBPA did not activate human and mouse PXR conversely to BPA which activated human PXR (Sui et al. 2012); however, by using transfected HG5LN-PXR cells, both TCBPA and TBBPA displayed agonistic activity similarly to BPA (Molina-Molina et al. 2013). In same cells, TCBPA,
<figure><img src="image_26.png" /></figure>

<figure><img src="image_27.png" /><figcaption>**Figure 1.** Pathways related to thyroid hormone (TH) metabolism disruption by BPA and alternatives. Regulation of TH synthesis starts with the production of TRH in the hypothalamus that activates TSH production in the pituitary gland. TSH is the main regulator of thyroid follicular cells through the TSHR that activates several nuclear transcription factor (PAX8, FOXE1, NKX2.1, and HHEX), that subsequently activate the expression of genes involved in thyroid differentiation (Fernández et al. 2015). Dietary iodide (I⁻) enters the follicular cell through NIS, localized at the basolateral plasma membrane (De la Vieja and Santisteban 2018; Koumarianou et al. 2022). Later I⁻ crosses the apical membrane, mainly through PENDRIN, into the colloid where TPO, using H₂O₂ produced by DUOX2, oxidizes the tyrosine residues of TG to form different iodine-containing compounds (MIT, DIT, T3, and T4) (Di Jeso and Arvan 2016). The amount of T4 synthesized is much higher than that of T3. Under metabolic requirement of TH, TG is endocytosed into the follicular cell and the iodine-compound is enzymatically released. DEHAL1 enzyme discharges I⁻ from inactive iodide-containing compounds (MIT and DIT) to be reincorporated in the TH synthesis process (De la Vieja and Santisteban 2018; Riesco-Eizaguirre et al. 2021). THs (T3 and T4) are transported to the bloodstream mainly through MCT8 where they are bound to TH transporters (TBG, TTR, and HAS) and delivered to different peripheral tissues (Mimoto and Refetoff 2020). TH is transported to peripheral cell cytoplasm by different transporters (MCT8, MCT10, LAT1, LAT2, OATP1C1, OATP2B1, and OATP3A1) in a tissue-specific manner (Bernal et al. 2015). Iodothyronine deiodinases enzymes DIO1 and DIO2 in cell cytoplasm remove one I⁻ from the inactive TH T4 to convert it to the active TH T3, that can be inactivated to T2 by removing an additional I⁻ by DIO3 (Darras and Van Herck 2012). T3 enters the nucleus where it activates the expression of a large number of genes in a tissue-specific manner by engaging TR and forming a complex with retinoid X receptor (RXR) and other co-activators/repressors that bind thyroid hormone response elements (TRE) located in the regulatory regions of target genes (Vella and Hollenberg 2017). THs also regulate their own synthesis by inhibiting the production of TRH and TSH in the hypothalamus and pituitary, respectively (Larsen 1982). THs enter the brain either directly via the blood-brain barrier (BBB) or indirectly via the blood-CSF-barrier (B-CSF-B) where MCT8 and OATP1C1 are essential (Bernal et al. 2015). The disruption of TH metabolism by BPA has been studied at different levels and in different *in vivo* and *in vitro* models. As can be seen in the top table, most of the genes/proteins essential in this process have been studied (with the exception of PENDRIN and DEHAL1). However, in the case of the remaining BPA analogues numerous gaps (blank table cells) remain. Furthermore, in many cases, an apparent lack of analogue effect (indicated as “NE?”) was observed, although this could be due to low tested concentrations (see article). TH: thyroid hormones; TRH: thyrotropin-releasing hormone; TSH: thyroid-stimulating hormone; TSHR: TSH receptor; PAX8: paired-box gene 8; FOXE1: forkhead box protein E1; NK2.1: NK homeobox 1; HHEX: hematopoietically expressed homeobox 1; NIS: sodium/iodide symporter; pendrin, anion exchange protein encoded by SLC6A4 gene; TPO: thyroid peroxidase; DUOX2: dual oxidase 2; TG: thyroglobulin; MIT: monoiodotyrosine; DIT: diiodotyrosine; DEHAL1: iodotyrosine dehalogenase 1; TBG: thyroxine-binding globulin; TTR: transthyretin; HAS: human serum albumin; MCT8/10: monocarboxylate transporter 8/10; LAT1/1: large neutral amino acids transporter small subunit 1/2; OATPs: organic amphipathic transporters; DIO1/2/3: type 1/2/3 iodothyronine deiodinase; TR: nuclear thyroid hormone receptor; NE: non-effect; ↑: increasing or upregulation; ↓: decreasing or downregulation; B↑/↓: biphasic up/down-regulation; The “?” symbol means that there are insufficient or contradictory results.</figcaption></figure>

BPZ, and BPP were strong PXR agonists, BPA, BPAP, and TBBPA were moderate agonists, whereas BPE was inactive (Grimaldi et al. 2019). Similarly, in transfected COS-7 cells, BPAP and BPZ were strong agonists, BPA and BPP were less potent and BPE had no activity (Kojima et al. 2019).

Concerning other receptors, the literature is very limited and thereby the following available data are not included in Table A1 and Table 2. Among the receptors involved in detoxification of xenobiotics, the aryl hydrocarbon receptor

(AhR) is particularly relevant, since it shares similar functionality with them, although, not belonging to the NR family (Mackowiak and Wang 2016). TBBPA was inactive in a DR-CALUX assay (Hamers et al. 2006), whereas in a reporter gene assay BPE displayed agonistic effect toward AhR with a potency higher than that of BPA (Rosenmai et al. 2014).

BPE, BPAP, and BPZ displayed a strong binding affinity for the constitutive androstane receptor (CAR), lower than BPA, whereas BPP was inactive (Liu X et al. 2019). No BPA
<figure><img src="image_28.png" /></figure>

alternative was a CAR agonist, rather all of them were antagonists with the following potency order BPZ > BPAP > BPA, BPP, and BPE being inactive (Kojima et al. 2019).

In competitive binding assays, BPA, BPAP, BPE, BPP, and BPZ were inert toward retinoic acid receptors (RARα, RARβ, and RARγ), and RAR-related orphan receptors (RORα, RORβ, and RORγ). Only BPP had very weak activity for retinoid X receptors (RXRα, RXRβ, and RXRγ) (Liu X et al. 2019). However, BPAP, BPE, BPZ, and BPP were previously shown not to bind RORγ, as BPA (Nishigori et al. 2012). In a reporter gene assay, TBBPA had no agonistic activity for RARα, but displayed a strong antagonism, higher than BPA (Xu H et al. 2022). In addition, TBBPA had no activity for hRXRα in transfected 3T3-L1 cells; however, RXRα gene expression was increased in same cells by TBBPA at 1 pM and 10 µM (Chappell et al. 2018). Conversely, TCBPA displayed antagonist activity for RXRβ over 7-fold higher than BPA (Li N et al. 2016).

BPA, BPAP, BPE, and BPZ were inert toward the liver X receptors (LXRα, LXRβ) as well as the vitamin D receptor (VDR), whereas BPP was inactive toward VDR but had strong binding affinity for both LXRs (Liu X et al. 2019). TCBPA and TBBPA exhibited a binding potency 8.7- and 34-fold higher than BPA, respectively, for the farnesoid X receptor (FXR). However, their transactivation activity was lower than BPA, significantly inducing FXR expression only at the highest concentrations 25-50 µM (Zhang et al. 2023).

No compound showed agonist activity toward the mineralocorticoid receptor (MR); conversely, BPZ, BPAP, BPE, BPP, TCBPA, and TBBPA were all strong antagonists comparable to BPA (Grimaldi et al. 2019).

Overall, the evidence related to these receptors is too limited to draw conclusions on the activity of BPA alternatives. Further, BPS-MAE was never investigated. More studies are thus needed to clearly assess their ED potential considering all the possible receptor pathways.

### 3.2.7. Concluding considerations

Overall, a lot of data gaps exists on the ED activity of the prioritized BPA alternatives, especially on BPS-MAE, still poorly investigated. Importantly, the reported results reflect only the activity of parent compounds since no information is available on the metabolic competence of the cell lines used and the activity of the potential metabolites, with consequent biases in the qualitative and quantitative aspects of effect estimation.

## 3.3. Effects on thyroid hormone system pathways

Even though there is ample evidence of potential TH system disruption by BPA (see Table A2), large data gaps remain regarding the potential effects of BPA alternatives (Table 3). In rats, BPA effects on different parameters of the thyroid function have been more consistently associated with high doses (>40 mg/kg/bw per day), which are far above the highest estimated BPA human dietary intake (infants and toddlers up to 0.875 µg/kg/bw per day; adult men and women up to 0.388 µg/kg/bw per day; adolescents up to 1.449 µg/kg/bw per day), even after applying a HED factor of 0.068, according to EFSA criteria (EFSA 2015). All levels of regulation of the thyroid system and/or thyroid-related physiological processes are potential targets of bisphenols (Figure 1, Table 3, and Table A2). In this respect, TH receptors are discussed here in the context of TH system disruption.

### 3.3.1. Effects on TH levels

Animal studies linking BPA exposure to TH levels have shown variable results, with TH levels increasing (Zoeller et al. 2005; da Silva et al. 2019), decreasing (Viguié et al. 2013; Ahmed 2016; Jiang et al. 2016; Guignard et al. 2017; Silva BS et al. 2019; Mohammed et al. 2020), or not changing (Kobayashi et al. 2005; Sadowski et al. 2014; Bansal and Zoeller 2019). In relation to this variability, studies showed biphasic responses and/or responses depending on sex, age, exposure route, and duration (Xu X et al. 2007; Santos-Silva et al. 2018). Among the BPA alternatives into focus in this review, data on TH levels are available only for TBBPA. All the related studies (Van der Ven et al. 2008; Saegusa et al. 2009; EFSA 2011; Cope et al. 2015; Osimitz et al. 2016; Sanders et al. 2016), but one (Meerts et al. 1999), consistently showed a decrease in T4 concentrations. Results on T3 or thyroid-stimulating hormone (TSH) are less consistent.

The paucity of basic data on the effect of BPA alternatives on TH levels in vivo highlights the huge gaps and needs for data to characterize the effects and modes of action of these molecules on the thyroid system.

### 3.3.2. Effects on thyroid hormone receptors (THRs)

BPA is a weak ligand of THRs, inhibiting T3-induced response in *Thr*α and *Thr*β gene reporter assays in a dose-dependent manner (Moriyama et al. 2002). A structure-based study indicated that BPA analogues, including BPP, BPZ, and BPAP could form important polar and hydrophobic interactions with several THRα residues, most of which also bind the native ligand T₃ (Beg and Sheikh 2020). In a fluorescence competitive binding assay, TBBPA displayed relative binding potencies to THRα and β about 15- and 100-fold lower than that of T₃, respectively (Ren et al. 2020).

Different gene reporter assays evidenced modest transactivation potency for THRα of TBBPA, BPA, and BPE from 10 to 60 µM, 30 and 70 µM for TBBPA, BPA, and BPE, respectively i.e. much lower than that of T₃, while TCBPA showed low or no activity (Jugan et al. 2007; Lei et al. 2017). However, in a yeast two-hybrid assay incorporating rat Thrα, TBBPA, and TCBPA displayed agonist activity (Terasaki et al. 2011). In rat GH3 cell line, BPA, BPAP, BPP, BPZ, TBBPA, and TCBPA all showed a positive effect on GH3 cell proliferation suggestive of THR agonism (Kitamura et al. 2002; Lee J et al. 2018). For TBBPA and TCBPA, however, antagonistic effects on T3-mediated responses were also shown (Jugan et al. 2007).

Neither BPA nor BPAF (10⁻⁹ to 10⁻⁵ M) appeared to influence *Thr* expression in transfected human granulosa cell line COV434 (Mlynarcikova and Scsukova 2020). By contrast, in rat GH3 cell lines, BPA and BPAP decreased *Thr*α and *Thr*β expression at 44 and 34 µM, respectively (Lee S et al. 2017).
<figure><img src="image_29.png" /></figure>

<figure><img src="image_30.png" /><figcaption>**Figure 2.** Schematic representation of the available evidence on effects exerted by BPA alternatives on steroidogenesis from *in vitro* studies in human H295R cells and mouse MA10 Leydig cells. Detailed data are provided in Table A3. Colored boxes indicate induction (orange), inhibition (blue), or inactivity (grey). White boxes indicate that the compounds were not assessed. Steroidogenic enzymes are indicated in blue italics. CYP: cytochrome P450; HSD: hydroxysteroid hydrogenase.</figcaption></figure>

Furthermore, BPP had no effect on *Thr*α expression within the range of 30 nM to 30 µM, while it inhibited *Thr*β expression from 3 µM in the same cell line. BPZ decreased *Thr*α and *Thr*β expression in a concentration-dependent manner from 0.37 to 3.7 µM for *Thr*α and *Thr*β, respectively (Lee S et al. 2017).

BPA, at low concentrations (nM), might negatively regulate THR transcriptional activity via an indirect mechanism involving the recruitment of nuclear receptor corepressor (N-CoR) or silencing mediator for retinoid and thyroid-hormone receptor (SMRT) to THRβ1 (Moriyama et al. 2002; Sheng et al. 2012). No data are currently available suggesting that other bisphenols might trigger such mechanisms.

Overall, evidence on halogenated derivatives of BPA suggests they may exert some agonistic effects on THRs. BPE showed some transactivation activity only toward THRα and indirect evidence of some agonism is available for BPAP and BPP in relation to cell proliferation. However, data on gene expression modulation of THRs mostly indicate an inhibitory effect of these compounds; therefore, more studies are needed to clarify this contrasting evidence. In addition, to our knowledge, information is lacking for BPS-MAE.

### 3.3.3. Effects on TH biosynthesis

*In vivo* data, mostly obtained with high to very high doses (20 up to 400 mg/kg/d), are suggestive of inhibitory effects of BPA and/or TBBPA on TH biosynthesis whatever the evaluated parameter: thyroid histology, gene expression, iodine uptake, and TPO activity (da Silva et al. 2018; Mohammed

et al. 2020; Hu C et al. 2023). In FRTL5 rat thyroid cells, BPA (100 µM) decreased *Slc5a5* and *Tpo* while increasing *Tg* (from 30 µM), *Pax8* and *Foxe1* (100 µM) expressions (Wu Y et al. 2016). This was correlated with a noncompetitive inhibition of iodine capture by BPA at high concentration (100 µM) while no impact on TPO activity could be shown (Wu Y et al. 2016). Interestingly, in FRTL-5 cells treated with BPA, a biphasic response was observed for *Pax8* and *Foxe1* expression (Gentilcore et al. 2013). BPP, BPZ, or TBBPA had no effect on *Dio1*, *Dio2*, *Tshβ*, *Tshr*, *Slc5a5*, *Nkx2.1*, *Pax8*, *Tg*, and *Tpo* expressions (Lee S et al. 2017). In this same study, BPA and BPAP seemed to positively regulate the transcription of *Slc5a5* at quite high concentrations of 0.4 mM and 34 µM for BPA and BPAP, respectively. Interestingly, in PCCL3 rat thyroid cell, BPA at low concentration (10 nM) decreased *Tpo* and *Slc5a5* mRNA levels (da Silva et al. 2018). BPA can inhibit *Tshβ* expression in adult female rats at very high dose (da Silva et al. 2018) and in GH3 cells at high concentration (Lee S et al. 2017). BPAP and BPZ had no effect on *Tshβ* expression in GH3 cells (Lee S et al. 2017).

In conclusion, most of the BPA alternatives studied seemed to not affect TH biosynthesis pathways, except TBBPA, which decreased the expression of some related genes in mice and BPAP, which was able to induce *Slc5a5* gene expression in an *in vitro* model. No data are available for TCBPA, BPE, and BPS-MAE; thus, further studies are warranted to complete the information on all the prioritized BPA alternatives.
<figure><img src="image_31.png" /></figure>

### 3.3.4. Interaction with TH serum transporters

Some studies showed that BPA can bind to serum TH transporters such as human serum albumin (HSA), thyroxine-binding globulin (TBG), and transthyretin (TTR) (Cao et al. 2011), although this seemed to be concentration-dependent, with little binding activity in the nM range (Meerts et al. 2000; Marchesini et al. 2008). All data on BPA binding capacities to TTR agreed on showing a very low potency of BPA to displace TH from TTR (Yamauchi et al. 2003; Kudo et al. 2006; Šauer et al. 2021). By contrast, available data indicate that TBBPA and TCBPA exhibited much higher binding capacities (Yamauchi et al. 2003; Kudo et al. 2006; Šauer et al. 2021). In particular, TBBPA had a relative potency about 10-fold higher than that of T4 to displace labeled-T4 from TTR, whereas TCBPA was very similar to T4 (Meerts et al. 2000). BPE behaved similarly to BPA toward T4 binding to TTR (Šauer et al. 2021). A bioinformatics study showed that BPP and BPAP should be able to form polar and hydrophobic interaction with the native T4 binding residues in TBG (77–100%) (Beg and Sheikh 2020).

Overall, although some evidence exists on the binding activity of some BPA alternatives to TH serum transporters no experimental reports related to TTR binding are available for BPZ, BPP, BPAP, and BPS-MAE. In addition, information is lacking for the complete panel of prioritized BPA alternatives as regards binding to HSA and TBG.

### 3.3.5. Interaction with TH plasma membrane transporters

The transport of TH into or out of cells is critical in TH-related processes, not only in the thyroid, but also in all target tissues and especially at the level of the placenta during pregnancy and in the brain in stages of embryonic and fetal development (Bernal et al. 2015). An impact of BPA on MCT8-mediated T3 uptake was shown on MDCK cells overexpressing human MCT8 (Dong and Wade 2017) but at very high concentration (250 µM) at the limit of cytotoxicity. This was not confirmed at lower concentrations on primary mouse astrocytes or MCT-8 transfected cells (Johannes et al. 2016). To the best of our knowledge, there are no available data evaluating the effect of BPZ, BPE, BPP, BPAP, PBS-MAE, TCBPA, and TBBPA on MCT8-mediated TH uptake.

### 3.3.6. Concluding considerations

To summarize, from the review of the numerous experimental data available on the effects of BPA on the thyroid-related pathways, it can be legitimately hypothesized that BPA is not neutral on thyroid regulation. For BPA alternatives, because of the paucity of available data (for summary see Table 3), many gaps remain that should be addressed in a well-structured frame with validated assays and with dose ranges relevant to human exposure. Our analysis highlights in particular the lack of "functional endpoints" such NIS or TPO activities.

## 3.4. Effects on steroidogenesis

Compared to the studies investigating interaction of BPA alternatives with NRs, effects on steroidogenesis are

somewhat underexplored. Available evidence is detailed in Table A3 and summarized in Figure 2.

Only two reports (Rosenmai et al. 2014; Lin et al. 2021) analyzed impairment of steroid hormone synthesis in the human adreno-carcinoma cell line H295R as per the OECD Test Guideline 456 (OECD 2023) performing complete dose-response assessments (Rosenmai et al. 2014), or evaluating effects of chemicals at three selected concentrations (Lin et al. 2021).

BPE induced both estrone and E2 secretion in H295R cells with EC₅₀ values comparable to those of BPA (Rosenmai et al. 2014). While BPA did not affect progesterone secretion and decreased 17α-OH progesterone levels, BPE induced both progesterone and 17α-OH progesterone. This would support a different mechanism for BPA and BPE to affect E2 synthesis, even if both compounds decreased dehydroepiandrosterone (DHEA) and androstenedione, upstream the synthesis of sex steroids. Both BPA and BPE strongly reduced testosterone synthesis in H295R cells, with comparable potency. BPA decreased cortisol levels, in agreement with 17α-OH progesterone drop, and had no effect on corticosterone. On the contrary, BPE did not affect cortisol levels but increased corticosterone level, according to 17α-OH progesterone increase (Rosenmai et al. 2014).

BPP, BPZ, and BPAP did not affect E2 secretion in H295R cells, conversely to BPA (Lin et al. 2021). BPA significantly decreased testosterone secretion at 25 nM whereas BPAP decreased it at both 6.25 and 25 nM and BPP did not affect hormone levels. A non-monotonic dose-response was observed for BPZ, which induced a testosterone drop at 0.39 and 6.25 nM but not at 1.56 nM (Lin et al. 2021).

TBBPA did not affect E2 and testosterone levels at 0.025, 0.05, and 0.5 µM in H295R cells (Song et al. 2008). In MA-10 mouse Leydig cells, BPA decreased E2 secretion at 1 nM; similarly, TCBPA 1 nM and TBBPA 10 nM decreased the secretion of E2, TCBPA to a higher extent (Gorowska-Wojtowicz et al. 2019). In same cells, estrone but not E2 was assessed, being unaffected by TBBPA at 10 µM, which otherwise increased both progesterone and 17α-OH progesterone levels. TBBPA, at same concentration, also increased the synthesis of 11-deoxycorticosterone, androstenedione and testosterone. For testosterone, a dose-response increase was also observed, being significant at 30 µM and 100 µM, with a maximum 59-fold increase compared to control cells (Roelofs et al. 2015). A similar dose-response in MA-10 cells had been previously observed, with a 46-fold increase for TBBPA 30 µM, compared to the sixfold induction by BPA (Dankers et al. 2013). E2 secretion was time- and dose-dependently increased by TBBPA in JEG-3 human choriocarcinoma placental cells. However, these cells are not able to synthesize E2 from cholesterol, thus effects were assessed in presence of the precursor DHEA (Honkisz and Wójtowicz 2015a). TBBPA also increased progesterone levels in JEG-3 cells with a similar time- and dose-dependency (Honkisz and Wójtowicz 2015b). In human placental explants, TBBPA significantly increased testosterone secretion and dose-dependently induced progesterone secretion, although with marginal significance only at the highest dose (range 5 nM to 50 µM). Only at 5 µM, a
<figure><img src="image_32.png" /></figure>

significant drop in E2 production was reported (Arita et al. 2018), in contrast to what observed in JEG-3 cells.

Beside steroid hormones, TBBPA significantly decreased the human chorionic gonadotropin (β-hCG) production in JEG-3 cells at all time points (24-72 h) (Honkisz and Wójtowicz 2015b). Moreover, both TCBPA and TBBPA increased β-hCG protein expression in mouse MA-10 Leydig cells (Gorowska-Wojtowicz et al. 2019).

As regards enzymes involved in the steroidogenic pathway, BPZ significantly repressed *CYP11A* expression at 6.25 nM in H295R cells, and did not affect levels of *CYP17*, *CYP19*, steroidogenic acute regulatory protein (*StAR*) and 3β-hydroxysteroid dehydrogenase type 2 (3β-HSD2). BPAP had no effect on any enzyme (Lin et al. 2021). In same cells, TBBPA had no effect on aromatase activity (*CYP19*) (Cantón et al. 2005), but induced *CYP21* expression (Song et al. 2008). TBBPA significantly increased *StAR*, *Cyp11A1*, and *Cyp17* expression in one study (Dankers et al. 2013) whereas no effect was observed in another report (Roelofs et al. 2015). Both studies confirmed the lack of effects on 3β-Hsd1 and 17β-Hsd3 expression (Dankers et al. 2013; Roelofs et al. 2015). Conversely, in JEG-3 placental cells, TBBPA induced a time- and dose-response increase in aromatase enzyme activity and protein expression (Honkisz and Wójtowicz 2015a).

### 3.4.1. Concluding considerations

Overall, TBBPA was the most investigated analogue in relation to all hormones and genes of steroidogenesis. One report assessed BPE effects on all steps of steroidogenesis. TCBPA was assessed only for effects on E2 secretion in mouse cells, whereas BPP, BPZ, and BPAP were assessed only for effects on E2 and testosterone production and some steroidogenic enzyme expression. No evidence on BPS-MAE is available and thereby additional studies are needed to have comparable results for all BPA alternatives, especially using validated methods as the OECD Test Guideline 456.

# 4. Immunotoxicity of BPA alternatives

## 4.1. General considerations

Immunotoxicology studies the adverse effects on the immune system caused by exposure to various physical and chemical agents. This complex system includes immunocompetent cells and soluble factors with specialized roles, divided into specific adaptive immunity and nonspecific innate immunity. The characteristics of the immune system make it vulnerable to xenobiotics (Germolec et al. 2022). Damage to this system can be associated with a wide range of adverse and sometimes life-threatening events, including immunosuppression, inappropriate immunostimulation, hypersensitivity, and autoimmunity. These effects are typically assessed using *in vitro* assays and *in vivo* models in both laboratory animals and humans (Maddalon et al. 2023).

Once believed to work independently from each other, it is now clear that the endocrine and the immune systems are tightly interwoven (Manley et al. 2018). As a close link exists between the endocrine and the immune systems, and the

immune system is highly receptive to endocrine signals due to the expression of hormone receptors on immune cells, any endocrine active substance is likely to affect the immune response (Sabuz Vidal et al. 2021).

Increasing evidence indicates, especially following developmental exposure, an ability of BPA in possibly modulating immune responses and signaling pathways, which leads to a proinflammatory response by favoring the differential polarization of immune cells (e.g. decreased T regulatory cells, increased T helper (Th) 17 cells) and cytokine production profile to one consistent with proinflammation (Sharif et al. 2022). In the recent EFSA opinion on the reevaluation of the risks to public health related to the presence of BPA in foodstuffs, effect on Th17 cells in mice was identified as the critical effect (EFSA 2023); these cells are pivotal in cellular immune mechanisms and involved in the development of inflammatory conditions, including autoimmunity and lung inflammation. In addition, epidemiological studies have discussed potential associations between several autoimmune diseases and BPA exposure, including neuroinflammation in the context of multiple sclerosis, colitis in inflammatory bowel disease, nephritis in systemic lupus erythematosus, and insulitis in type 1 diabetes mellitus (Sharif et al. 2022). However, diverging views on BPA proinflammatory effects exist and should be considered (BfR 2023; BfR and EFSA 2023). In addition, more reports discuss an increased propensity to develop allergy, and asthma after dietary and inhalation exposure to BPA (Robinson and Miller 2015; McDonough et al. 2021).

## 4.2. Information on the immunotoxicity of BPA alternatives

While there are many original studies and reviews on the immunotoxicity of BPA (Robinson and Miller 2015; Xu J et al. 2016; Aljadeff et al. 2018; Sabuz Vidal et al. 2021; Sharif et al. 2022), only few studies have addressed the effects of BPA alternatives on the immune system (McDonough et al. 2021). In this section, a review of data regarding the ability of BPA alternatives to perturb the immune system is presented.

There is some evidence of BPP effects on the immune system. Indeed, mice exposed for 5 or 9 weeks to BPP (0.3, 30, and 3000 µg/kg bw/day) exhibited gut microbiota dysbiosis, together with the activation of lipopolysaccharide/nuclear factor-kappaB (NF-κB) pathway and inflammation (Ma N et al. 2023).

Effects on immune parameters were also investigated for BPAP. BPAP exposure (400 µg/kg bw/day) induced hippocampal inflammation, microglial inflammation, and a higher number of splenic macrophages and activated dendritic cells in mice (Wu X et al. 2023). These effects were associated with increased anxiety state level as reported in Section 5.

No studies evaluating the effects of BPZ or BPE exposure on the immune system are available. However, as described in Section 3, BPZ is threefold more potent than BPA in agonizing ERα and BPE has a similar potency than BPA in inducing ERα (Pelch KE et al. 2019). Given the tight connection between the endocrine and immune systems, it
<figure><img src="image_33.png" /><figcaption>**Table 4.** *In vitro* studies reporting neural effects of BPA alternatives (BPAP, BPE, BPP, BPZ, and TCBPA).</figcaption></figure>

<table><thead><tr><th>Compounds</th><th>Dose range</th><th>Cell types</th><th>Analyses</th><th>Reference</th></tr></thead><tbody><tr><td>BPA, BPAP, BPB, BPE, BPF, BPS, and BPZ</td><td>All substances at 1, 10, and 100 nM in 0.01% DMSO</td><td>(i) Human embryonic stem cells (hESCs)<br>(ii) hESC differentiated into neural epithelium<br>(iii) Neuron-like cells differentiated from adherent neuronal stem cells</td><td>Cytotoxicity severity was assessed for all targeted compounds based on the IC<sub>50</sub> calculation or estimation; BPAF was defined as the most cytotoxic chemical tested whereas BPF and BPS the less cytotoxic.<br>The results pointed out that BPA and its analogues did induce neurogenesis impairment (neural epithelium generation) at early stages of neurogenesis but did not exclude that the latter could interfere with a later stage of neurodevelopment. Gene expression related to axon guidance process was not modified by the exposure of BPA and its analogues; normalized values of the total and maximum- neurite length were down-regulated by BPA in a dose dependent manner.<br>A significant reduction of total neurite length was noted for 4 out of the 6 BPA analogues tested (BPS, BPF, BPA, or BPAF) at 1 nM; while only 100 nM of BPB, BPZ, or BPE triggered a noticeable decrease in the total length of neurite.</td><td>Liang X et al. (2020)</td></tr><tr><td>BPA, BPAP, BPBP, BPC, BPE, BPG, BPP, and BPPH</td><td>All substances at 1 nM, 100 nM and 10 µM for 3 h or 24 h.</td><td>IMR-32 neuroblastoma cell line</td><td>Treatment for 24 h:<br>All bisphenols except BPPH increased ROS levels, since 1 nM (BPA, BPE) or 100 nM (BPC) or at 10 µM (BPG, BPAP, BPP).<br>The protein levels of Bax, Bak1, and caspase-3 were increased by 100 nM except for BPP and BPPH. The apoptosis rate was also increased since 100 nM (BPA, BPC) or at 10 µM (BPE, BPG, BPP, BPBP).<br>BPA, BPE, BPC, BPG, and BPAP at 10 µM and BPAP at 100 nM also significantly increased Ca<sup>2+</sup> levels in IMR-32 cells.<br>Treatment for 3 h:<br>BPA, BPE, BPC, BPAP, and BPP at 10 µM increased intracellular Ca<sup>2+</sup> levels in IMR-32 cells, suggesting an activation of GPER.<br>This effect was attenuated by GPER antagonist (G15) treatment (except for BPP). The ROS levels induced by BPC and BPAP were also reduced by G15.<br>Molecular dynamics simulation suggests that BPA, BPE, BPC, and BPAP may activate GPER.</td><td>Wang L et al. (2023)</td></tr><tr><td>TCBPA, TBBPA, TBBPS, BDE-47, and BDE-209</td><td>TCBPA, TBBPA, TBBPS, BDE-47, as well as mix of 5 compounds: 0.01, 0.1, 1, and 5 µM<br>BDE-209 doses: 10, 100, and 1000 nM</td><td>Human embryonic stem cell (HESC)</td><td>Deregulation of transcription factors determinant for neuronal development like ZIC1, ZIC3, HES3, IGFBP3 as well as DLX5 by the five targeted chemicals.<br>All the 5 chemicals targeted dysregulated genes (e.g. CNTN2, SLIT1, LRRC4C, RELN, CBLN1, CHRN84, and GDF7) involved in axon growth/guidance and neuron transmission-related processes.<br>AhR and WNT signaling pathways were impacted by all compounds. All five flame retardants significantly up-regulated CYP1A1 gene expression, suggesting that all of them could behave as agonists for the AhR signaling pathway.</td><td>Liang S et al. (2019)</td></tr></tbody></table>

Ach: acetylcholine; AMPAR: α-amino-3-hydroxy-5-methyl-4-isoxazolepropionic acid receptor; AhR: aryl hydrocarbon receptor; Bak1: BCL2 antagonist/killer 1; Bax: Bcl-2-associated X; BFR: brominated flame retardant; CBLN1: cerebellin 1 precursor; CGC: cerebellar granule cells; CHRN84: cholinergic receptor nicotinic beta 4 subunit; CNTN2: contactin 2; DLX5: distal-less homeobox 5; GDF7: growth differentiation factor 7; GPER: G protein-coupled estrogen receptor; HES3: Hes family BHLH transcription factor 3; HFR: halogenated flame retardant; hESCs: human embryonic stem cells; IGFBP3: insulin-like growth factor binding protein-3; LOEC: the lowest observed effect concentration; LRRC4C: leucine rich repeat containing 4C; NLRP3: NOD-, LRR-, and pyrin domain-containing protein 3; NMDAR: N-methyl-D-aspartate receptor; RELN: Reelin; ROS: reactive oxygen species; SLIT1: slit guidance ligand 1; ZIC: zinc finger protein.

The full version including also data on TBBPA is available in Table A5.

is reasonable to hypothesize an immune adverse action of both bisphenols. BPE short-term exposure (24 h) was also able to induce oxidative stress, increasing the level of ROS

on neuroblast cells at the concentration of 1 nM, 100 nM, and 10 µM (Wang L et al. 2023). Also in this case, the existing connection between oxidative stress and inflammation,
<figure><img src="image_34.png" /></figure>

renders presumably an effect of BPZ and BPE on immunity (Lauridsen 2019).

There are relatively more data available for halogenated bisphenols such as TBBPA. This bisphenol was able to reduce tumor necrosis factor alpha (TNF-α) levels in both peripheral blood mononuclear cells (PBMCs) and purified natural killer (NK) cells starting at the lowest tested concentration of 0.05–5 µM (Yasmin and Whalen 2018). Similar effects were also observed on interferon gamma (IFN-γ) on human PBMC, monocyte-depleted PBMCs and purified NK cells, following 24 h, 48 h, or 6 days of exposure at the same concentrations active on TNF-α (Almughamsi and Whalen 2016). Instead, the secretion of interleukin (IL)-1β was increased in the same cells by the same concentrations of TBBPA 10 and 20 µM (Anisuzzaman and Whalen 2016). In trophoblasts, TBBPA increased the release of IL-6 and IL-8 and reduced the release of transforming growth factor beta (TGF-β) (Park HR et al. 2014). NK lytic functionality was also reduced by TBBPA exposure starting from 0.5 µM (Kibakaya et al. 2009). Furthermore, TBBPA has been linked to the downregulation of several genes linked to immune system at the dose of 250 mg/kg bw/day, suggesting an immunosuppressive effect in vivo (NTP 2013; Hall et al. 2017). In addition, TBBPA (10–100 nM) was able to affect gene transcription on mouse embryonic stem cells, at two main levels: neuronal differentiation and immune functionality, with a reduction of the membrane T cell receptor Cd79a, Cd79b, interleukin 7 receptor (IL7r), colony-stimulating factor 1 receptor (Csf1r), C-C motif chemokine receptor 2 (Ccr2), and an increase of the high affinity IgE receptor (Fcer1g) genes (Tribondeau et al. 2022).

Exposure of female mice to TCBPA (5 and 50 mg/kg bw/day) resulted in immune perturbations. In detail, it increased both pro- and anti-inflammatory cytokine levels in serum, namely IL-2, IL-4, IL-5, IL-10, IL-12, TNF-α, IFN-γ, and granulocyte-macrophage colony-stimulating factor (GM-CSF) (Wang Y et al. 2021). Exposure of mice for 14 days to the same TCBPA doses was able to decrease the percentage of T lymphocytes but to increase T regulatory cells, and the highest concentrations of TCBPA (100 µg/mL) also included a reduced cell proliferation in mouse primary splenic lymphocytes exposed to TCBPA for 72 h (Wang Y et al. 2021).

### 4.3. Concluding considerations

To date, there are limited available data on the immunotoxic potential of BPA alternatives. However, there is an indication of possible effects on the immune system for some of them, affecting both innate and acquired immunity, that could in turn contribute to several immune-mediated conditions. Additional studies are necessary to identify the most suitable BPA substitutes, also taking into account the metabolic competence of the *in vitro* system used.

## 5. Developmental neurotoxicity of BPA alternatives

### 5.1. General considerations

Neural effects have contributed to the classification of BPA as a SVHC (ECHA 2016, 2017). These effects included impaired

spatial and non-spatial memory and modifications in neural processes related to synaptic plasticity (Mhaouty-Kodja et al. 2018). Developmental exposure to other bisphenols such as BPS also induced behavioral effects in rodents (Beausoleil et al. 2022). Whether BPA alternatives such as BPZ, BPE, BPP, BPAP, BPS-MAE or TBBPA and TCBPA have adverse effects comparable to BPA is discussed below. The retrieved *in vivo* and *in vitro* data are presented in Table 4 and in Tables A4 and A5, respectively.

### 5.2. Neurotoxic effects of BPA analogues and halogenated bisphenols

#### 5.2.1. BPA analogues

A recent study assessed the potential neurodevelopmental effects of BPAP (Wu X et al. 2023) but no data are currently available for BPZ, BPE, BPP, or BPS-MAE. Prenatal and postnatal exposure of mice to 0.4 mg/kg/d of BPAP reduced the novel object recognition behavior, sociability, and social novelty behavior in adult males and females (Table A5). In addition, BPAP-exposed animals exhibited increased anxiety state level in the open-field, elevated plus maze, and marble burying tests. These behavioral modifications were associated with changes in hippocampal gene expression. In particular, genes associated with neurodevelopment including neural cell adhesion molecule 2 (Ncam2) and interleukin 1 receptor accessory protein like 1 (Il1rap1) were down-regulated. In contrast, microglial and neuroinflammatory markers were upregulated, suggesting microglial activation that was confirmed at the protein level by increased immunoreactivity of ionized calcium-binding adapter molecule 1 in the hippocampus of BPAP-exposed mice.

An *in vitro* study examined the impact of BPA and its analogues BPS, BPF, BPE, BPB, BPZ, and BPAF on cell differentiation processes (Table 4). As human embryonic stem cells (hESCs) and hESC-derived neural stem cells (NSCs) can be differentiated into neural epithelium and neuron-like cells, respectively, they have been used to investigate the effects of bisphenol exposure during differentiation on gene and protein expression levels as well as on cell morphology (Liang X et al. 2020). BPAF was the most cytotoxic analogue in both cell models (0 up to 300 µM). BPS, BPF, BPA, and BPAF all triggered morphological changes, with a significant reduction in neurite total length from the lowest dose tested (1 nM), while a slight decrease was shown for BPE, BPB, or BPZ at the highest dose tested (100 nM). When cells were exposed at concentrations that mimic a human exposure scenario (1–10 nM), no change was observed in the levels of expression of genes and proteins involved in the process of neural epithelium from hESC or specification of neuron-like cells from NSCs (Liang X et al. 2020).

A more recent study evidenced the potential of BPA and several of its alternatives, including BPP and BPE, to induce oxidative stress and apoptotic processes in a neuroblastoma cell line (Wang L et al. 2023). This effect involved mechanisms mediated by GPER, recognized as a major mediator of non-genomic estrogen signaling pathways (Alexander et al. 2017). Unlike the effects of bisphenols mediated by the
<figure><img src="image_35.png" /></figure>

nuclear ERs, these effects occurred at relatively low concentrations, within the nM range for increased ROS levels with BPA and BPE. Although it increased caspase 3 expression by a factor of two, BPAP appeared to have the weakest effect on ROS expression and did not show any effect on apoptosis rate. The effect of BPA, BPE, and BPAP but not those of BPP on calcium mobilization and ROS formation were reversed by the GPER antagonist G15 highlighting the endocrine mediation of these adverse effects of bisphenols on this neural cell line (Wang L et al. 2023).

### 5.2.2. Halogenated bisphenols

Concerning the halogenated bisphenols, given that no *in vivo* data were available on the neurodevelopmental effects of TCBPA, the review focused on the effects of TBBPA. Hence, we analyzed the nine *in vivo* studies addressing the effects of prenatal/postnatal and prepubertal/pubertal exposure to TBBPA, as presented in Table A4.

In eight studies, the effects of prenatal/postnatal exposure to TBBPA were demonstrated at the doses indicated in Table A4. These included effects on the auditory response in rat, with suggested cochlear effect in females and a more apparent neural effect in males (Lilienthal et al. 2008). The benchmark doses of auditory effects were comparable to those reducing TH levels, and lower than those reducing maternal body weight. Perinatal exposure of rats to TBBPA at 0.1 mg/kg/d increased wheel activity in females, and elevated anxiety-state level in males after continuous exposure to TBBPA at 25 and 250 mg/kg/d (Rock et al. 2019). Postnatal exposure to a TBBPA bis(2-hydroxyethyl ether) (TBBPA-BHEE), a representative TBBPA derivative, also triggered motor coordination and activity deficits in prepubertal male rats at the two doses used of 0.086 and 0.86 mg/kg (Liu QS et al. 2018). Acute postnatal exposure of male mice to TBBPA increased activity in the open-field test at 5 mg/kg, freezing in the contextual fear conditioning paradigm at 0.1 and 5 mg/kg, and spontaneous alternation in the Y maze at 0.1 mg/kg (Nakajima et al. 2009). Deficits in memory retention using the passive avoidance test (Wang J and Dai 2022) and in spatial memory using the Morris water maze (Kim AH et al. 2017) have been also described following prepubertal/pubertal exposure of male mice to TBBPA at doses above 50 mg/kg/d. Finally, a reduction in social interaction was induced in adult male rats exposed prenatally and postnatally to TBBPA at 0.2 mg/kg/d (Kim B et al. 2015).

At the neuroanatomical level, changes have been reported in the number of hippocampal interneurons in the developing brain (Saegusa et al. 2012). Alterations in gene expression profile analyzed by microarray in rats exposed to TBBPA-BHEE (Liu QS et al. 2018), as well as in hippocampal neurogenesis (Kim AH et al. 2017), oxidative stress, apoptosis markers, brain-derived neurotrophic factor (BDNF), and neurotransmitter levels in mice have also been described (Wang J and Dai 2022). It has been suggested that TBBPA-induced neurodevelopmental effects involved direct effects on the brain (Saegusa et al. 2012) and indirect actions through modifications in TH (Lilienthal et al. 2008). When male mice were exposed to a single postnatal TBBPA dose of

11.5 mg/kg, no changes were observed in the levels of proteins related to hippocampal neuroplasticity (Viberg and Eriksson 2011).

In one study assessing the effects of developmental exposure to TBBPA on the F2 generation of rats born from F1 parents exposed prenatally to 10–1000 mg/kg/d of TBBPA, Cope et al. (2015) reported no detectable neurodevelopmental or neurobehavioral deficits.

In addition to the *in vivo* studies, there are *in vitro* studies mostly focused on the neurotoxic effects of TBBPA (see Table A5). Exposure to TBBPA triggered oxidative stress, intracellular Ca²⁺ imbalance, elevation of extracellular glutamate and death in primary culture of cerebellar granule cells (CGCs) prepared from seven-day old Wistar rats (Reistad et al. 2007; Zieminska et al. 2017; Bowen et al. 2020). The increase in intracellular [Ca²⁺], oxidative stress, and mitochondria depolarization was dose-dependent from 10 to 25 µM (Al-Mousa and Michelangeli 2012; Zieminska et al. 2016; Diamandakis et al. 2019). The authors hypothesized that microsomal Ca-ATPase inhibition could play a crucial role in Ca²⁺ homeostasis, which in turn induces mitochondrial dysfunction (Al-Mousa and Michelangeli 2012). In similar conditions of treatment, TBBPA also impaired the homeostasis of Zn²⁺ in neuronal cells (Zieminska et al. 2017; Bowen et al. 2020). Acute exposure of CGC to TBBPA (10 or 25 µM) triggered a steady decline in cell viability (from 3 h up to 24 h) associated with an increasing number of neurons with irregularities in their nuclei. At the highest dose, TBBPA altered gene expression involved in the major signaling pathways that lead to neuronal death. A decrease in the expression of autophagy and anti-apoptotic genes was observed in the first 3 h of exposure, followed, in a second step, by the activation of apoptosis-associated genes from 6 to 24 h post exposure. Expression of programmed necrosis-associated genes was not altered regardless of the exposure time (Lenart et al. 2017). Similar results were obtained on primary hippocampal neuronal cultures prepared from Swiss mouse embryo at gestational days 17–18. Exposure to TBBPA at doses ranging from 100 nM to 100 µM similarly led to apoptotic cell death, although in these cells it was accompanied by a certain level of necrotic cell death (Szychowski and Wójtowicz 2016). The ability of TBBPA to induce cell death, at doses ranging from 3 to 10 µM, at least partially by apoptosis, through the activation of caspase has also been described in SH-SY5Y human neuroblastoma cells at low µM levels (Al-Mousa and Michelangeli 2012). The authors pointed out a clear relationship between the alteration of Ca²⁺ homeostasis within neuronal cells and the initiation of neuronal cell degeneration and death. Furthermore, the damage done by 12 h of TBBPA exposure increased secretion of β-amyloid-42 levels, a key molecular factor in neurotoxicity (Al-Mousa and Michelangeli 2012) and expression of the presynaptic synaptosomal-associated protein, 25 kDa (SNAP-25) that controls neurotransmitter release and plasma membrane growth (Zieminska et al. 2016). These results suggest that SNAP-25 might be a target leading to cognitive ability impairments. Ototoxicity was induced by exceptionally high TBBPA levels (125 µg/mL) in HEI-OC1 (Park C et al. 2016). This observation agrees with the significant increase in the percentage of apoptotic cells in Corti organ explants from postnatal day 2 rats exposed to TBBPA for 18–
<figure><img src="image_36.png" /><figcaption>Table 5. Summary of studies addressing the prioritized BPA alternatives' genotoxicity.</figcaption></figure>

<table><thead><tr><th>Substance</th><th>Cell model</th><th>Assay</th><th>Exposure conditions</th><th>Effect</th><th>Reference</th></tr></thead><tbody><tr><td rowspan="4">BPAP</td><td rowspan="2">HepG2 3D spheroids</td><td>Comet assay</td><td>0.01, 0.1, 1, 10, and 40 µM, for 24 h<br>0.01, 0.1, 1, and 10 µM, for 96 h</td><td>Positive at the highest dose<br>Positive at ≥0.1 µM</td><td rowspan="2">Sendra et al. (2023)</td></tr><tr><td>γH2AX</td><td>0.01, 0.1, 1, 10, and 40 µM, for 24 h<br>0.01, 0.1, 1, and 10 µM, for 96 h</td><td>Negative<br>Positive</td></tr><tr><td rowspan="2">RAD54<sup>-/-</sup> mutant chicken DT40 cells</td><td>γH2AX</td><td>31.3 µM, for 1 h</td><td>Negative</td><td rowspan="2">Lee S et al. (2013)</td></tr><tr><td>Chromosomal aberration</td><td>7.8, 15.6, and 31.3 µM, for 48 h</td><td>Negative</td></tr><tr><td rowspan="2">BPP</td><td rowspan="2">RAD54<sup>-/-</sup> mutant chicken DT40 cells</td><td>γH2AX</td><td>12.5 µM, for 1 h</td><td>Positive</td><td rowspan="2">Lee S et al. (2013)</td></tr><tr><td>Chromosomal aberration</td><td>3.13, 6.25, and 12.5 µM, for 48 h</td><td>Positive at the highest dose</td></tr><tr><td rowspan="2">BPZ</td><td rowspan="2">HepG2 cells</td><td>Comet assay</td><td>0.1, 1, and 10 µM, for 4 and 24 h</td><td>Positive at ≥1 µM, for 4 h<br>No cytotoxicity at 12.5, 25, 50, and 100 µM, for 24 h</td><td rowspan="2">Fic et al. (2013)</td></tr><tr><td>Ames test</td><td>4, 20, 100, and 500 µg/plate ± S9</td><td>Negative</td></tr></tbody></table>

γH2AX: histone family member X; HepG2 cells: human hepatoma-derived cell line; RAD54<sup>-/-</sup>: defective in RAD54 gene; S. typhimurium: Salmonella typhimurium.

24 h (Park C et al. 2016). However, the levels of exposure used to observe these neuro- and ototoxic effects are far from being representative of human exposure levels, which would be more in the µM or even nM range (Liang X et al. 2020). The capacity of TBBPA to induce mitochondrial bioenergetic impairments was also investigated in murine BV-2 microglial cells. BV-2 cells remained viable (≥80% of viability) following exposure to TBBPA concentrations ranging from 10 to 40 µM (Bowen et al. 2020). However, although weak changes in pro-inflammatory responses were observed, there were significant alterations in mitochondrial energy production, including ATP-linked respiration, basal, and compensatory glycolysis. The phagocytotic potential was also decreased (Bowen et al. 2020).

Only one *in vitro* study investigated TCBPA-induced effects (Liang S et al. 2019) (Table 4). The neurotoxic effects of TBBPA alone or in combination with other flame retardants (BDE-47, BDE-209, TBBPS, and TCBPA) were investigated on induced neural ectoderm from hESCs (Liang S et al. 2019). RNA-seq analyses highlighted that all the halogenated flame retardants (HFRs) targeted triggered dysregulation in the expression of genes involved in (i) transcription factors determinant for neural development, (ii) axon growth/guidance, and (iii) neuron transmission processes. Assays conducted on HFR mixture did not show clear synergic effects (Liang S et al. 2019). Based on the obtained results, Liang S et al. (2019) concluded that, from a neurotoxic point of view, TCBPA should not be considered as safe alternative to TBBPA.

### 5.3. Concluding considerations

The SVHC classification of BPA supported the plausibility that the impaired learning and memory and underlying synaptic plasticity processes induced by this substance were essentially mediated through disruption of the estrogenic signaling pathway (Mhaouty-Kodja et al. 2018). Major endocrine

systems such as the steroid and thyroid pathways are well-known regulators of neural development and cognitive function (Hamson et al. 2016; Mughal et al. 2018). Alterations of these pathways whether at the maternal or the fetal levels during critical windows of neurodevelopment might result in long-term effects. To our knowledge, there are no *in vivo* data linking neurodevelopmental effects of BPA alternatives to an endocrine mode of action. The available *in vitro* data presented in Section 3 together with the study of Wang J and Dai (2022) suggest the ability of some bisphenol alternatives to bind and activate receptors for sex steroid and THs, but more studies are still required to confirm this hypothesis.

Overall, the scarce data reported for BPA alternatives (one *in vivo* study for BPAP and two *in vitro* studies for several BPA analogues) together with the available data on TBBPA suggest that exposure to BPA alternatives might induce potential neurodevelopmental effects. However, additional studies are required to address these aspects. In addition, although DNT with an endocrine component was one main endpoint that led to the classification of BPA as a SVHC at the EU level, data gaps prevent a proper evaluation of the BPA alternatives/analogues effects on DNT via endocrine-mediated mechanisms.

## 6. Genotoxicity and carcinogenicity of BPA alternatives

### 6.1. General considerations

The characterization of the carcinogenic potential of BPA alternatives is central for their hazard and risk assessment. For the classification of chemicals' carcinogenicity, the International Agency for Research on Cancer (IARC) considers all relevant and reliable data contributing to the knowledge on carcinogenicity in experimental animals, and mechanistic evidence from *in vitro* and *in vivo* models (IARC 2019).
<figure><img src="image_37.png" /></figure>

Regarding the mechanistic evidence, Smith et al. (2016) proposed the application of the 10 key characteristics of human carcinogens (e.g. genotoxicity, induction of cell proliferation, immunosuppression or epigenetic effects, among others) to systematically identify and organize mechanistic information about a given agent. Through this approach, Ricker et al. (2024) concluded that BPA may act by multiple and interconnected mechanisms. Indeed, exposure to BPA has been associated with breast, ovarian, prostate, cervical, and lung cancer development, possibly due to its ED activity (Seachrist et al. 2016; Almeida et al. 2018; Khan et al. 2021; Winz and Suh 2021). Several rodent studies showed that BPA influences the action of estrogen and progestin agonists through interactions with ERs, GPER, and PR, leading to mammary tissue hyperplasia and cancer development (Vandenberg et al. 2008; Perrot-Applanat et al. 2018), which contributed to the BPA classification as a SVHC (ECHA 2017). The assessment of BPA genotoxicity through *in vitro* testing showed no mutagenic or clastogenic potential in some studies (Ivett et al. 1989; Schweikl et al. 1998; Tsutsui et al. 1998; Audebert et al. 2011; Fic et al. 2013) while others evidenced its ability to produce DNA and chromosomal damage, particularly in human liver-derived cell lines (Fic et al. 2013; Sendra et al. 2023; Yu H and Liu 2023) and in ER-positive MCF-7 cells (Iso et al. 2006). Overall, based on the available data, EFSA concluded that it is unlikely to very unlikely that BPA presents a genotoxic hazard by a direct mechanism of action (EFSA 2023).

## 6.2. Information on the genotoxicity and carcinogenicity of BPA alternatives

Recent studies have supported the effects of data-rich BPA alternatives (e.g. BPF and BPS) and other less studied alternatives, including TCBPA, on cell proliferation and cell migration through an ER-dependent pathway (Lei et al. 2018; Awada et al. 2019; Huang et al. 2019; Jung et al. 2019; Lei, Tang, et al. 2021; Lei, Xu, et al. 2021; Yu M et al. 2023). This suggests that those chemicals cannot be considered a priori safe alternatives for BPA replacement (Edaes and de Souza 2022).

Regarding the genotoxicity of the prioritized BPA alternatives, Table 5 summarizes the main knowledge available in the literature. A recent *in vitro* study showed that BPAP (0.01–10 µM, 96 h exposure) induces the formation of DNA and chromosome breaks (using the comet and the γH2AX assays, respectively) in human liver-derived HepG2 3D spheroids (Sendra et al. 2023). However, no clastogenicity had been previously observed in RAD54⁻/⁻ mutant chicken DT40 cells (Lee S et al. 2013). BPP showed genotoxic potential by inducing chromosomal aberrations (12.5 µM, 1 h) and γH2AX foci formation (12.5 µM, 48 h) in the referred DT40 cells (Lee S et al. 2013). BPZ was able to induce DNA strand breaks (1 and 10 µM, 4 h exposure) in HepG2 cells. BPZ, similarly to other BPA alternatives (BPF, BPAF, and BPS) was not mutagenic in the Ames test, with and without exogenous S9 metabolic fraction (Fic et al. 2013). For the remaining prioritized BPA alternatives, i.e. BPE, BPS-MAE, and TCBPA, no published data on their potential genotoxicity were found, which does not exclude the existence of datasets, e.g. as parts of

chemicals' request for authorization by competent authorities.

On the other hand, there is information on other BPA alternatives' genotoxicity and the data available, particularly, regarding BPS, BPF, and BPAF, are also briefly presented due to their relevance to understand possible mechanisms of action of the prioritized analogues (see Table A6). BPF was shown to be genotoxic in HepG2 cells (Fic et al. 2013; Lee S et al. 2013; Hercog et al. 2019; Yu H and Liu 2023). Likewise, BPS was mostly genotoxic by inducing DNA breaks and/or chromosomal alterations in HepG2 (Fic et al. 2013; Sendra et al. 2023; Yu H and Liu 2023), human prostatic (Kose et al. 2020), and bronchial BEAS-2B cells (George and Rupasinghe 2018), and also in a humanized peripheral blood mononuclear cell (Hu-PBMC) model (Mokra et al. 2017). Several studies have also evidenced BPAF genotoxicity, especially showing a clastogenic effect (Lee S et al. 2013; Mokra et al. 2017; Hercog et al. 2019; Sendra et al. 2023). Mokra et al. (2017) observed that BPAF and BPS were the substances bearing the highest and the lowest genotoxic potential in Hu-PBMCs, respectively. BPAF was the most potent inducer of oxidative DNA damage, whereas BPS induced the lowest level, suggesting that oxidative damage was implicated in the observed genotoxicity (Mokra et al. 2018). The authors associated the ability of bisphenols to induce oxidative DNA damage to their possible intracellular biotransformation into highly reactive hydroxylated metabolites, which could be further transformed into quinones via CYP 2E1, existent in mammalian lymphocytes (Mokra et al. 2018). The importance of the *in vitro* cell systems' metabolic competence to reliably assess the genotoxicity of BPA alternatives was demonstrated by Yu H et al. (2020). Using Chinese hamster lung-derived cell lines expressing various human CYP enzymes and a human (metabolism-proficient) hepatoma C3A cell line, the authors showed that BPF and BPS (40, 80, and 160 µM, for 9 h followed by a 15 h of recovery), like BPA, were clastogenic (Yu H et al. 2020). Consistent with those observations, Hercog et al. (2019) reported that BPF, BPS, and BPAF upregulated the expression of CYP1A1 in HepG2 cells. Moreover, they observed a negative clastogenicity modulation by phase II enzymes (SULTs and UGTs) (Yu H et al. 2020) suggesting that inactive metabolites are formed. Audebert et al. (2011) had previously shown that BPF was extensively metabolized by HepG2 and LS174T intestine cells and produced a genotoxic effect.

Although some studies have suggested that bisphenols increase intracellular ROS through an estrogen-mediated pathway (Lei et al. 2018), ER-independent oxidative effects and chromosomal alterations have also been observed (Špačková et al. 2020). It is therefore plausible that the genotoxicity of these substances may be mediated by (ER-dependent or -independent) oxidative stress (Ribeiro et al. 2017). Epigenetic effects such as DNA methylation, histone modification, and the expression of noncoding RNAs are some of the epigenetic targets of BPA and its alternatives (reviewed by Almeida et al. 2018) that might explain an estrogen-independent genotoxic and carcinogenic mode of action. However, the link between DNA damage caused by bisphenols exposure and epigenetic effects remains unclear.
<figure><img src="image_38.png" /></figure>

## 6.3. Concluding considerations

Overall, the knowledge about potential genotoxic effects of the selected BPA alternatives is very limited. Although some evidence exists for the *in vitro* genotoxicity of BPAP, BPP, and BPZ, no published data were found for the other prioritized chemicals. Given the ability of other more data-rich BPA alternatives (BPS, BPF, and BPAF) to induce genotoxic effects, possibly mediated by oxidative stress after undergoing phase I metabolism, there is an urgent need for expanding the knowledge on the BPA alternatives in focus. Low-concentration ranges relevant for human exposure should be tested under standardized conditions, using metabolism-proficient cell lines or external metabolic systems to account for the potential effect of metabolites. In addition, considering the key characteristics of carcinogens, other features, such as the ability of altering cells proliferation, inducing oxidative stress, or cell transformation, among others, should be investigated.

## 7. Conclusions, knowledge gaps, and future directions

Although manufacturers have already started using alternative chemicals to BPA for many applications, there are a number of features which should be considered to avoid regrettable substitutions of BPA in particular, and of any chemical of concern, in general. Among those features, which include performance (similar functional properties), bio-degradability, cost-effectiveness, transparency, and innovation, those related to safety and regulatory compliance are of utmost importance. While more data on potential adverse effects of already marketed BPA alternatives are needed, those under development should undergo rigorous safety assessment. Ideally, they should have a proven track record of safety through comprehensive testing, possibly using NAMs. This means that BPA should be replaced by nontoxic chemicals, both in their raw form and when used in consumer products. This includes ensuring that they do not leach harmful chemicals into food, beverages, or the environment, even under conditions such as heating or acidic environments. Despite the data already produced on their biological activity, most alternatives to BPA still present a major gap of knowledge regarding their metabolism and kinetics, as well as effects on endocrine and immune systems, neurodevelopment, genotoxic and carcinogenic properties, as detailed in the concluding considerations at the end of each section or subsection.

The generation of kinetic data has been identified as one of the key research priorities needed to improve mechanistic understanding of chemical-induced toxicity (European Commission 2012; EFSA 2014; Escher et al. 2022), and the knowledge on BPA kinetic behavior has been crucial to decrease the uncertainties related to its risk assessment based on toxicity data produced in animal models (EFSA 2010, 2015, 2023). On the other hand, when NAMs are used and integrated in the NGRA, kinetics data are essential as PBK input and for developing a robust QIVIVE. Despite these considerations, kinetic data are available for a few analogues, such as BPS and BPF, while scant, fragmentary or no

information could be found for the alternatives selected within the PARC project.

The review highlighted that the reported *in vitro* toxicity data are generally obtained with cell lines endowed with no, scant, or unknown metabolic competence, thus assessing the activity of the parent compound. This may represent a substantial source of overestimation or underestimation of the overall hazard, depending on whether metabolism generates inactive or bioactive metabolites, respectively. Therefore, the use of metabolism-proficient *in vitro* cell models becomes crucial for accurate dose-response identification with these models. In addition, as previously discussed, the concentrations tested *in vitro* are frequently much higher with respect to the possible human exposure levels indicated by biomonitoring studies (in the nM to low µM range). The extrapolation to low dose can be another relevant source of uncertainty.

In summary, it will be important to study the potential mode(s) of action of BPA alternatives, in order to establish their ED potency relative to that of BPA, and to obtain comparable data using the same cellular systems and harmonized methodologies. In addition, data still need to be generated concerning their DNT or immunotoxicity, considering that the neural effects were among the endpoints leading to the classification of BPA as a SVHC at the EU level (ECHA 2017), and immunotoxic effects were used for the health-based reference value in the most recent EFSA opinion on BPA risk assessment (EFSA 2023). Furthermore, despite the scarcity of data, there are studies showing that some BPA alternatives own key characteristics of carcinogens, including the potential to produce genotoxicity, epigenome alterations, and cell proliferation, especially in estrogen-insensitive tissues, such as the liver. Therefore, there is a clear need to also produce data on these endpoints and underlying mechanisms, in order to clarify whether the BPA alternatives in use might contribute to the development of cancer mediated (or not) by a genotoxic mechanism.

The data available and reviewed here also evidenced that, despite a lot of similarities with BPA, the different alternatives also show a number of differences in their kinetics and induced effects. A good example is given by the wide spectrum of estrogenic activities, with TBBPA being almost inactive, while others display a different level of activity toward the various ERs with no information available on the biotransformation products. The same considerations can be applied to the other receptor binding activities and, as a consequence, the potentially mediated toxic effects. This implies that in most cases the read across principle with data on BPA or on other data-rich BPA analogues (i.e. BPS or BPF) should be used with caution, possibly based on robust bridging studies obtained with the different BPA alternatives.

The reported data originated from an extensive literature search (although the papers were not retrieved following the strict principles of a systematic review) carried out by experts in the field of each endpoint, in order to obtain a level of detail appropriate for the aim of the PARC project and the target audience. In addition, the main focus was on BPA alternatives relevant for the European Regulatory Agencies, as the major stakeholders within the frame of the PARC project.
<figure><img src="image_39.png" /></figure>

Some of the conclusions drawn in the present review are shared by another recent review also related to PARC activities that summarized, from the environmental perspective, the gaps and needs on BPA alternatives (Adamovsky et al. 2024). In agreement with the present analysis, nearly 70% of studies using vertebrate models focused on BPS, BPF, BPAF, and TBBPA. Adamovsky et al. (2024) stressed the importance of conducting more comprehensive assessments that go beyond the traditional reproductive studies and focus on overlooked relevant endpoints, such as the immune system. Future research should also consider mixture effects, realistic environmental concentrations, and the long-term consequences on biota and ecosystems.

Overall, from the present review we conclude that there is an urgent need to (i) determine the mechanism(s) through which BPA alternatives (and/or their metabolites) act, (ii) develop relevant in vitro tools that allow assessing simultaneously ED and other adverse outcomes, and (iii) search for molecular biomarkers that allow predicting immunotoxicity, endocrine disruption, non-genotoxic carcinogenesis, and/or DNT of novel BPA alternatives, before they come into widespread use. A step forward should also include the evaluation of sex-specific effects due to sexual dimorphism of many of the endpoints herein discussed. Moreover, the future in vitro studies should (i) test concentration ranges close to the actual human exposure, (ii) include, whenever possible, biokinetic measurements showing intracellular uptake and distribution, and (iii) have the potential to study the effects of metabolites by using metabolism-competent cells.

Within PARC, the majority of identified gaps will be addressed and studying the key steps of BPA alternatives' mode(s) of action will be a crucial process underlying a tight synergism with other internal projects to improve/build AOPs, including also kinetics data. This is relevant for all endpoints selected, crucial to classify or not the selected substances as EDCs, as well as to determine if they are safe alternatives to BPA by carrying out an adequate quantitative risk assessment.

## Acknowledgements

The authors would like to express their sincere gratitude to the anonymous external reviewers selected by the Editor for their thorough review, comments, and suggested revisions, which greatly helped improving the final manuscript.

## Declaration of interest

This review was prepared on the initiative of a small group of authors (editorial group), who agreed on its general structure and the information to be included. The editorial group worked together with the remaining coauthors to develop the content of the various sections, draft the text and revise it. The group was also responsible for reading, harmonizing and revising the text, to get the final submitted version. All authors developed the work in the course of their employment at their respective organizations and under the scope of the Partnership for the Assessment of Risks from Chemicals (PARC) that has received funding from the European Union's Horizon Europe Research and Innovation Programme under Grant Agreement No. 101057014. None of the authors declared a competing financial interest; none declared professional or personal relationships that could have appeared to influence the work

presented or constitute a conflict of interest. The European Partnership for the Assessment of Risks from Chemicals 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 Programme under Grant Agreement No. 857560. The authors also acknowledge co-funding by their organizations and the following entities: Centre for Toxicogenomics and Human Health (ToxOmics), DOI 10.54499/UIDP/00009/2020 (https://doi.org/10.54499/UIDP/00009/2020) and DOI 10.54499/UIDB/00009/2020 (https://doi.org/10.54499/UIDB/00009/2020); Spanish Ministerio de Ciencia e Innovación (MICIN) and Fondo Europeo de Desarrollo Regional (FEDER) PID2021-125948OB-I00/AEI/10.13039/501100011033. Dr. Adamovsky thanks The Czech Science Foundation (GACR), Project 23-13617L, for support in plastic and plastic additives research.

## ORCID

Emanuela Testai

<figure><img src="image_40.png" /></figure>

http://orcid.org/0000-0003-3113-5103

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