Bisphenols: Endocrine Disruptors and Their Impact on Fish: A Review

Abstract

Bisphenols (BPs), particularly bisphenol A (BPA) and its structural analogues, are synthetic compounds widely used in plastics and industrial materials. These substances are also recognised as endocrine-disrupting chemicals (EDCs) due to their ability to interfere with hormonal systems, which has significant implications for aquatic organisms. This review summarises the occurrence, environmental distribution, and toxicity of BPs in fish, with a focus on estrogenic, androgenic, thyroid, and glucocorticoid disruptions. Studies consistently show that exposure to BPs leads to altered gene expression, developmental abnormalities, impaired reproduction, and disrupted hormonal signalling in various fish species. Although BPA alternatives like bisphenol S, bisphenol F, or bisphenol AF were introduced as safer options, emerging evidence suggests they may pose equal or greater risks. Regulatory measures are evolving, particularly within the European Union, but legislation remains limited for many bisphenol analogues. This review emphasises the need for comprehensive environmental monitoring, stricter regulatory frameworks, and the development of genuinely safer alternatives to minimise the ecological and health impacts of BPs in aquatic systems.

1. Introduction

In recent decades, a significant and growing threat to fishes and other aquatic organisms has emerged due to the increasing global human population and the subsequent release of a wide variety of chemical wastes into freshwater ecosystems [1].

One particularly concerning group of chemicals are endocrine-disrupting compounds (EDCs), which can interfere with the normal functioning of the endocrine system in organisms. These chemicals, whether of natural or anthropogenic origin, are exogenous substances that interfere with the endocrine system, potentially mimicking or blocking hormones or altering the secretion, mechanisms of action, or rates of metabolism of endogenous hormones, thereby leading to adverse health outcomes [2]. In particular, they can lead to various types of disruption in organisms, such as estrogenic, androgenic, and thyroid disruption, among others. These compounds can be found in a wide array of sources, including industrial chemicals primarily used in the production of various plastics and epoxy resins, as well as pesticides and phthalates that have the potential to leach into food and water supplies [3,4]. Furthermore, even some pharmaceuticals and substances in personal care products can also act as EDCs and negatively impact the health and welfare of fish [5,6]. Bisphenols (BPs) are one such class of pollutants that have garnered increased attention due to their presence in the environment and their effects, as they have been linked to various adverse health effects in fish, including impaired reproduction, growth, and development [3,4].

BPs, a class of synthetic organic compounds, have gained significant attention recently because of their widespread presence in aquatic environments [7]. They can cause a wide range of physiological, behavioural, and developmental changes [8]. This review will explore in detail the endocrine effects of BPs on fish health and fitness.

2. Characteristics of Bisphenols and Their Occurrence in the Aquatic Environment

BPs are widely utilised in the production of a diverse array of products, including plastics, water pipes, toys, medical equipment, electronics, food cans, and various household applications [9,10]. As the name implies, BPs contain two phenol groups and serve as fundamental building blocks for polymers [11]. When exposed to acidic conditions or high temperatures, the ester bonds within the polycarbonate polymer can undergo breakdown, leading to the release of BPs into contained foods or beverages [11].

Among the most significant BPs, bisphenol A (BPA) stands out due to its widespread use and confirmed adverse effects, particularly its potential as an EDC. BPA has been extensively detected in various environmental compartments, leading to its regulation or ban in certain products [8,12]. As a result, efforts have been made to find substitutes for BPA. However, concerns exist that these alternatives may exhibit similar toxicological profiles. Some of the most common structural analogues include bisphenol S (BPS), bisphenol F (BPF), and bisphenol B (BPB), with others numbering in the tens [13]—a summary of the most commonly used BPA and particular bisphenol analogues, which are further mentioned, is given in

Table 1. BPA and other specific bisphenol analogues, which are most commonly used and further discussed in the review, along with their chemical formula and structure.

Bisphenol Type	Chemical Formula	Structural formula
Bisphenol A (BPA)	C15H16O2
Bisphenol S (BPS)	C12H10O4S
Bisphenol F (BPF)	C13H12O2
Bisphenol AF (BPAF)	C15H10F6O2
Bisphenol B (BPB)	C16H18O2
Bisphenol E (BPE)	C14H14O2
Bisphenol Z (BPZ)	C18H22O2
Tetrabromobisphenol A (TBBPA)	C15H12Br4O2

.

Bisphenol analogues, used as substitutes for BPA, have significant environmental impacts. They are frequently detected in aquatic environments and can affect marine species by disrupting the neuroendocrine system, causing oxidative stress, and leading to developmental alterations [10,14]. These analogues, such as BPF and BPS, exhibit endocrine-disrupting effects similar to or greater than BPA, impacting estrogen and androgen receptor signalling in aquatic organisms [15]. Additionally, bisphenol analogues can influence soil health by affecting soil microbiomes and enzyme activity, which are crucial for plant growth and development [16]. The persistence and fate of these compounds in the environment remain areas of concern, as they can lead to long-term ecological effects [17].

Their environmental prevalence and presence in aquatic systems depend on their concentrations, which can vary significantly across different regions and ecosystems [11]. BPs have been detected in surface waters, sediments, and aquatic organism (biota) samples collected from various locations worldwide, indicating their widespread environmental distribution. Concentrations of BPs in the aquatic environment typically range from a few nanograms per litre to several micrograms per litre in surface waters, from a few nanograms per gram to hundreds of nanograms per gram in sediments, and from a few nanograms per gram to tens of nanograms per gram in biota, depending on the specific location and source of contamination [9,18].

The release of BPs into the environment can occur through various pathways, such as the breakdown of plastic products, industrial waste, and wastewater effluents [9,19]. They occur in nature not just due to their presence in wastewater but also in wastewater treatment plants, agricultural fertilisers, and thermal paper, resulting in widespread environmental contamination. This broad exposure from various sources underlines the necessity for continued research into their health effects and sources of intoxication [20]. The characteristics of BPs in the aquatic environment are of particular concern, as they can persist and accumulate in water bodies, sediments, and biota, thereby potentially impacting aquatic ecosystems [21]. Table 2 gives an overview of the occurrence of the most important representatives of BPs in various matrices of the aquatic ecosystem.

Table 2. Overview of the leading representatives of bisphenols and their occurrence in the aquatic environment (d.w. = dry weight; n/a = no data available; n.d. = not detected; w.w. = wet weight).

Bisphenol Type	Continent	Water	Fish (ng/g w.w.)	Sediment (ng/g d.w.)
Bisphenol A (BPA)	Asia	5.26–76.6 ng/L [22,23]; mean 23 ng/L [24]; <810 pg/L [25]	Coilia mystus: ~4–6, Pseudorasbora parva: ~12–14, Cyprinus carpio: ~11–13, Silurus asotus: ~11–14, [26]	0.56–5.22 [22]; mean 13.0 [24]; <0.6 [25]; 0.18–2010 [26]
	North America	<90.0 ng/L	n/a	<25,300 [18]
Bisphenol S (BPS)	Asia	0.07–5.2 ng/L [22,23]; 2.2 ng/L [24]	Coilia mystus: ~2–3, Pseudorasbora parva: ~1–2, Cyprinus carpio: ~1–2, Silurus asotus: ~1–2, [26]	n.d. −0.19 [22]; 0.69 [24]
Bisphenol F (BPF)	Asia	n.d.–12.6 ng/L [23]	Coilia mystus: ~2–3, Pseudorasbora parva: ~1–2, Cyprinus carpio: ~1–2, Silurus asotus: ~1–2, [26]	1.6 [24]
Bisphenol AF (BPAF)	Asia	0.44–10.8 ng/L [22,23]; 0.9–246 ng/L [26]	Coilia mystus: ~21, Pseudorasbora parva: ~13–17, Cyprinus carpio: ~3–4, Silurus asotus: ~8, [26]	0.08–0.66 [22]; 0.18–2010 [26]
Bisphenol B (BPB)	Asia	n.d.–14.3 ng/L [23]	Coilia mystus: ~2, Pseudorasbora parva: ~1–4, Cyprinus carpio: ~1–2, Silurus asotus: ~3, [26]	n/a
Bisphenol E (BPE)	Asia	n.d.–6.2 ng/L [23]	Coilia mystus: ~1–2, Pseudorasbora parva: ~1–2, Cyprinus carpio: ~1–2, Silurus asotus: ~1–2, [26]	n/a
Bisphenol Z (BPZ)	Asia	n/a	Coilia mystus: ~2–4, Pseudorasbora parva: ~2–3, Cyprinus carpio: ~1–3, Silurus asotus: ~1–3, [26]	n/a
Tetrabromo bisphenol A (TBBPA)	Asia	2.3 ng/L [24]; <810 pg/L [25]	n/a	<0.6 [25]

Table 2. Overview of the leading representatives of bisphenols and their occurrence in the aquatic environment (d.w. = dry weight; n/a = no data available; n.d. = not detected; w.w. = wet weight).

Bisphenol Type	Continent	Water	Fish (ng/g w.w.)	Sediment (ng/g d.w.)
Bisphenol A (BPA)	Asia	5.26–76.6 ng/L [22,23]; mean 23 ng/L [24]; <810 pg/L [25]	Coilia mystus: ~4–6, Pseudorasbora parva: ~12–14, Cyprinus carpio: ~11–13, Silurus asotus: ~11–14, [26]	0.56–5.22 [22]; mean 13.0 [24]; <0.6 [25]; 0.18–2010 [26]
	North America	<90.0 ng/L	n/a	<25,300 [18]
Bisphenol S (BPS)	Asia	0.07–5.2 ng/L [22,23]; 2.2 ng/L [24]	Coilia mystus: ~2–3, Pseudorasbora parva: ~1–2, Cyprinus carpio: ~1–2, Silurus asotus: ~1–2, [26]	n.d. −0.19 [22]; 0.69 [24]
Bisphenol F (BPF)	Asia	n.d.–12.6 ng/L [23]	Coilia mystus: ~2–3, Pseudorasbora parva: ~1–2, Cyprinus carpio: ~1–2, Silurus asotus: ~1–2, [26]	1.6 [24]
Bisphenol AF (BPAF)	Asia	0.44–10.8 ng/L [22,23]; 0.9–246 ng/L [26]	Coilia mystus: ~21, Pseudorasbora parva: ~13–17, Cyprinus carpio: ~3–4, Silurus asotus: ~8, [26]	0.08–0.66 [22]; 0.18–2010 [26]
Bisphenol B (BPB)	Asia	n.d.–14.3 ng/L [23]	Coilia mystus: ~2, Pseudorasbora parva: ~1–4, Cyprinus carpio: ~1–2, Silurus asotus: ~3, [26]	n/a
Bisphenol E (BPE)	Asia	n.d.–6.2 ng/L [23]	Coilia mystus: ~1–2, Pseudorasbora parva: ~1–2, Cyprinus carpio: ~1–2, Silurus asotus: ~1–2, [26]	n/a
Bisphenol Z (BPZ)	Asia	n/a	Coilia mystus: ~2–4, Pseudorasbora parva: ~2–3, Cyprinus carpio: ~1–3, Silurus asotus: ~1–3, [26]	n/a
Tetrabromo bisphenol A (TBBPA)	Asia	2.3 ng/L [24]; <810 pg/L [25]	n/a	<0.6 [25]

3. Regulatory Framework and Restriction of Bisphenol Usage

In the face of accumulating evidence on the adverse effects of BPA, regulatory agencies worldwide have taken steps to limit its usage in consumer products, particularly those intended for children and infants. These actions represent a crucial step towards mitigating the potential harm that BPA exposure can cause, particularly in vulnerable populations such as developing fetuses and infants [27,28]. However, as new research continues to uncover the far-reaching implications of BPA exposure, regulatory agencies must remain vigilant in monitoring its presence in consumer products and adjust their guidelines accordingly to safeguard public health [28].

The regulatory landscape surrounding BPA has evolved in response to mounting scientific evidence of its endocrine-disrupting effects and potential health implications [29]. Globally, regulations for controlling EDCs vary. The European Union (EU) has introduced regulations such as REACH, which focuses on a hazard-based approach, while the United States Environmental Protection Agency (EPA) implements a risk-based Endocrine Disruptor Screening Program (EDSP) [30,31]. In the USA, BPA has been banned in baby bottles since 2012 [32] and was banned in packaging for infants a year later [33], but it can be used in other food contact materials. Both approaches aim to identify and regulate harmful EDCs, although challenges remain in testing methods and distinguishing endocrine-specific effects [34]. BPA is also banned in baby bottles and other infant products in China [35], Canada [36], as well as South Korea [37] and other countries. Other analogues are not often covered yet, although monitoring is increasing.

The European Union has not imposed a complete ban on BPA but has progressively restricted its use due to its endocrine-disrupting properties and reproductive toxicity. Several legislative measures were initiated in 2011, when the EU banned the use of BPA in the manufacture of polycarbonate infant feeding bottles [38]. This decision was based on concerns about infants’ vulnerability to BPA exposure. Later in 2016, BPA was added to the Candidate List of Substances of Very High Concern (SVHC) under the REACH Regulation (EC) No 1907/2006 [39], due to its classification as a substance toxic for reproduction (category 1B). In 2017, the European Chemicals Agency (ECHA) further identified BPA as an endocrine disruptor for human health under REACH. Consequently, BPA was also classified as an endocrine disruptor for the environment in 2018. Commission Regulation (EU) 2016/2235 then prohibited the use of BPA in thermal paper (such as cash register receipts), with the restriction entering into force on 2 January 2020 [40].

The European Commission has also banned BPA in food contact materials due to its harmful health effects, based on EFSA’s latest assessment from 2023. The ban covers adhesives, rubbers, plastics, coatings, and other materials. Products made with other BPs or bisphenol derivatives must not contain residual BPA. Exceptions exist for polysulfone filtration membranes and liquid epoxy-based coatings in large-capacity food containers (>1000 L), provided that BPA migration into food is undetectable and the materials are properly cleaned before use. The regulation also prohibits other hazardous BPs affecting the reproductive and endocrine systems. The tolerable daily intake (TDI) for BPA has been drastically lowered to 0.2 ng/kg body weight, making even minimal migration unsafe. The current specific migration limit (SML) of 0.05 mg/kg is being phased out. The European Union Reference Laboratory (EURL) will develop methods to verify the absence of BPA, with a detection limit of 1 μg/kg, unless otherwise specified in Commission Regulation (EU) 2024/3190 [41]. Unfortunately, other analogues do not have unified legislation regulation due to the definitions in the regulation, as they do not fit the definition of “bisphenol derivatives” in the regulation [41], as shown in Figure 1.

The introduction of structural analogues as substitutes for BPA has further complicated the regulatory landscape [8,12]. As these alternative BPs have been found to exhibit similar endocrine-disrupting properties, regulatory agencies must extend their oversight to these compounds as well, ensuring that the public is protected from the potential health risks associated with exposure to the broader class of BPs [19,27]. In 2021, the European Chemicals Agency (ECHA) added BPB to the Candidate List of Substances of Very High Concern (SVHC) under REACH as well as BPS. For now, with BPA, they are the only BPs included due to their endocrine-disrupting properties and the rest are monitored and analysed. Last on the list is TBBPA for its carcinogenic potential, also included in 2023. But none of the BPs is currently included in REACH Regulation (EC) No 1907/2006; Annex XIV: List of substances subject to authorisation, as some of them are just candidates for Annex XIV [39]. Researchers and industry professionals alike have turned their attention to non-bisphenol materials, including plant-based polymers and other bio-based compounds, to identify viable replacements that minimise or eliminate the risks associated with BPs [42].

4. Toxicity of Bisphenols to Fish

BPs have been reported to induce a variety of effects in fish, including altered gene expression and the activity of key enzymes involved in the metabolism and production of endogenous hormones, physiological changes, and behavioural modifications [6,43,44]. In addition, BPs have also been implicated in the induction of oxidative stress [9,45,46]. The resulting oxidative stress has been linked to a range of adverse health effects, including inflammation, cell dysfunction, and tissue damage. Prolonged exposure to bisphenol-induced oxidative stress can have far-reaching health consequences, potentially contributing to the development of chronic diseases such as cancer, neurological disorders, and cardiovascular problems [27,47]. Evidence shows that exposure to BPs, even at relatively low concentrations, can lead to significant physiological, behavioural, and fitness-related changes in fish, such as cognitive disorders [48]. The change can then occur in local tissue, organs, or by mutations [49]. Apart from hormonal disruption, BPs can cause various problems in organisms.

Recent studies have revealed that alternative BPs can also induce toxic, reproductive, and neuroendocrine effects similar to those observed with BPA [10,19]. Specifically, these studies have shown that BPS and BPF can disrupt hormone signalling and lead to adverse health outcomes, including developmental and reproductive issues, in a manner comparable to the well-documented effects of BPA [13]. Numerous studies have investigated the toxic effects of BPs, with both animal and human studies demonstrating the harmful impacts of these chemicals [9]. One of the most well-documented effects of BPs is their estrogenic activity, which can disrupt the normal function of the endocrine system and lead to feminisation of male fish, altered sexual development, and impaired reproductive success [4,21]. However, the endocrine-disrupting effects of BPs extend beyond estrogen signalling, encompassing the disruption of androgen [3,50,51,52], thyroid [12,53,54,55,56,57], and glucocorticoid hormone pathways [58,59,60,61,62]. Different concrete disrupting effects are discussed in separate sections.

4.1. Estrogenic Activity and Disruption

Estrogenic disruption is a significant environmental concern, particularly due to the widespread presence of EDCs such as BPs [63,64]. BPs disrupt endocrine function primarily by mimicking the structure and function of natural estrogens, binding to estrogen receptors (ERs), and activating the estrogen signalling pathway (see Table 3 and Table 4 for an overview of significant findings on the estrogenic effects of selected BPs in fish). This interference disrupts normal hormonal balance, leading to consequences such as feminisation, impaired reproductive success, and changes in population dynamics [65]. Studies have demonstrated that even low concentrations of BPs can interfere with the reproductive axis, leading to impaired gonadal development, altered spawning behaviour, and reduced embryo viability [3,66]. Studies have shown that exposing male fish to estrogen-mimicking BPs can induce the production of vitellogenin (VTG) and lead to the development of intersex characteristics, thereby directly impacting the reproductive success and fitness of these animals [66,67].

Table 3. Overview of significant findings on the estrogenic effects of selected bisphenols in fish (↑ = increase; ↓ = decrease; dpf = days post-fertilisation; ER = estrogen receptor; hpf = hours post-fertilisation; VTG = vitellogenin).

Bisphenol Type	Fish Species	Exposure Duration	Concentration	Effects Observed	References
BPA	Danio rerio	15 days	0.01, 0.1, 1 mg/L	↑ egg production, ↓ fertilisation rate, altered reproductive gene expression, ↓ DNA methylation in ovaries, impaired reproductive processes	[68]
		30 days	1, 10, 100 µg/L	↑ atretic follicles, altered estrogen receptor expression, impaired ovarian function, reduced reproductive fitness	[3]
	Danio rerio, embryo–larvae	96 hpf–6 dpf	1–200 mg/L	↑ VTG, induced estrogenic response in heart, liver, muscle, and fins (ER-dependent)	[50,69]
	Danio rerio, male	6 weeks	100 and 2000 µg/L	early reproductive feminisation, female-like lipid metabolism, gonad damage, feminisation	[70]
	Danio rerio, larvae	5–14 dpf	up to 2500 µg/L	estrogenic response in the heart, altered heart rhythm, ↓ heart rate	[71]
	Pimephales promelas	43, 71, 164 days	1–1, 280 µg/L	↑ VTG in males (≥160 μg/L), gonadal growth inhibition (≥640 μg/L), ↓ egg production	[72]
		4 days	16–1, 280 µg/L	↓ gonadal growth, altered sex cell types, ↓ egg production at high concentrations	[69]
	Oryzias melastigma	70 days	200 µg/L	follicular atresia, irregular oocytes, empty follicles, ↓ eggs laid, ↓ fertilisation rate	[73]
	Carassius auratus	7–90 days	0.2 and 20 µg/L	↑ VTG (liver, 60 days), ↑ aromatase and ERs, estrogenic effects	[74]

Table 3. Overview of significant findings on the estrogenic effects of selected bisphenols in fish (↑ = increase; ↓ = decrease; dpf = days post-fertilisation; ER = estrogen receptor; hpf = hours post-fertilisation; VTG = vitellogenin).

Bisphenol Type	Fish Species	Exposure Duration	Concentration	Effects Observed	References
BPA	Danio rerio	15 days	0.01, 0.1, 1 mg/L	↑ egg production, ↓ fertilisation rate, altered reproductive gene expression, ↓ DNA methylation in ovaries, impaired reproductive processes	[68]
		30 days	1, 10, 100 µg/L	↑ atretic follicles, altered estrogen receptor expression, impaired ovarian function, reduced reproductive fitness	[3]
	Danio rerio, embryo–larvae	96 hpf–6 dpf	1–200 mg/L	↑ VTG, induced estrogenic response in heart, liver, muscle, and fins (ER-dependent)	[50,69]
	Danio rerio, male	6 weeks	100 and 2000 µg/L	early reproductive feminisation, female-like lipid metabolism, gonad damage, feminisation	[70]
	Danio rerio, larvae	5–14 dpf	up to 2500 µg/L	estrogenic response in the heart, altered heart rhythm, ↓ heart rate	[71]
	Pimephales promelas	43, 71, 164 days	1–1, 280 µg/L	↑ VTG in males (≥160 μg/L), gonadal growth inhibition (≥640 μg/L), ↓ egg production	[72]
		4 days	16–1, 280 µg/L	↓ gonadal growth, altered sex cell types, ↓ egg production at high concentrations	[69]
	Oryzias melastigma	70 days	200 µg/L	follicular atresia, irregular oocytes, empty follicles, ↓ eggs laid, ↓ fertilisation rate	[73]
	Carassius auratus	7–90 days	0.2 and 20 µg/L	↑ VTG (liver, 60 days), ↑ aromatase and ERs, estrogenic effects	[74]

Table 4. Overview of significant findings on the estrogenic effects of selected bisphenols in fish (↑ = increase; ↓ = decrease; dpf = days post-fertilisation; ER = estrogen receptor; hpf = hours post-fertilisation; VTG = vitellogenin).

Bisphenol Type	Fish Species	Exposure Duration	Concentration	Effects Observed	References
BPF	Danio rerio	96 h (embryo–larva)	1–100 µg/L	induced estrogenic response, ↑ ER and aromatase gene activity, ↑ VTG, altered reproductive neuroendocrine genes	[19,50,69]
		long-term (duration not specified)	1–100 µg/L	↑ expression of reproductive neuroendocrine genes (kiss1, gnrh3, lhβ, fshβ), ↑ VTG, ER and aromatase activity	[19]
		96 hpf–6 dpf	1–200 mg/L	↑ VTG in heart, liver, muscle, and fins	[50,69]
		21 days	0.1 and 1 mg/L	↑ VTG (males) and estrogenic effects	[75]
	Oryzias melastigma	70 days	200 µg/L	follicular atresia, irregular oocytes, empty follicles, ↓ eggs laid, ↓ fertilisation rate	[73]
BPAF	Danio rerio	4 hpf–120 dpf	5, 25, 125 µg/l	↑ 17β-estradiol in females, ↓ egg fertilisation	[61]
		96 h (embryo–larva)	1–200 mg/L	most potent estrogenic response, ↑ ER and aromatase gene activity, ↑ VTG	[50,69]
		96 hpf	100 and 200 µg/L	delayed gonadal migration, ↓ germ cell progenitors, altered hormone receptor expression	[76]
		96 hpf–6 dpf	1–200 mg/L	most potent ↑ VTG in heart, liver, muscle, fins	[50,69]
	Oryzias melastigma	70 days	200 µg/L	most pronounced ↓ eggs laid, follicular atresia, irregular oocytes, empty follicles, ↓ fertilisation	[73]
BPB	Danio rerio	21 days	0.1 and 1 mg/L	↓ egg production, estrogenic-like activity similar to or greater than BPA	[77]
		96 h (embryo–larva)	1–100 µg/L	↑ ER and aromatase gene activity, altered reproductive neuroendocrine genes	[78]
BPS	Danio rerio	4 hpf–120 dpf	1 and 100 µg/L	female-dominant sex ratio, impaired reproductive capacity	[61]
		75 days	0.1, 1, 10, 100 μg/L	↑ female to male sex ratio, ↑ estradiol and VTG (males, females), ↓ egg production	[79]
	Danio rerio, male	96 hpf–6 dpf	1–200 mg/L	↑ VTG at high concentrations, induced estrogenic response, less potent than BPA	[50,69]
		21 days	8, 40, 200 μg/mL	↑ VTG, ↑ aromatase and estradiol	[52]

Table 4. Overview of significant findings on the estrogenic effects of selected bisphenols in fish (↑ = increase; ↓ = decrease; dpf = days post-fertilisation; ER = estrogen receptor; hpf = hours post-fertilisation; VTG = vitellogenin).

Bisphenol Type	Fish Species	Exposure Duration	Concentration	Effects Observed	References
BPF	Danio rerio	96 h (embryo–larva)	1–100 µg/L	induced estrogenic response, ↑ ER and aromatase gene activity, ↑ VTG, altered reproductive neuroendocrine genes	[19,50,69]
		long-term (duration not specified)	1–100 µg/L	↑ expression of reproductive neuroendocrine genes (kiss1, gnrh3, lhβ, fshβ), ↑ VTG, ER and aromatase activity	[19]
		96 hpf–6 dpf	1–200 mg/L	↑ VTG in heart, liver, muscle, and fins	[50,69]
		21 days	0.1 and 1 mg/L	↑ VTG (males) and estrogenic effects	[75]
	Oryzias melastigma	70 days	200 µg/L	follicular atresia, irregular oocytes, empty follicles, ↓ eggs laid, ↓ fertilisation rate	[73]
BPAF	Danio rerio	4 hpf–120 dpf	5, 25, 125 µg/l	↑ 17β-estradiol in females, ↓ egg fertilisation	[61]
		96 h (embryo–larva)	1–200 mg/L	most potent estrogenic response, ↑ ER and aromatase gene activity, ↑ VTG	[50,69]
		96 hpf	100 and 200 µg/L	delayed gonadal migration, ↓ germ cell progenitors, altered hormone receptor expression	[76]
		96 hpf–6 dpf	1–200 mg/L	most potent ↑ VTG in heart, liver, muscle, fins	[50,69]
	Oryzias melastigma	70 days	200 µg/L	most pronounced ↓ eggs laid, follicular atresia, irregular oocytes, empty follicles, ↓ fertilisation	[73]
BPB	Danio rerio	21 days	0.1 and 1 mg/L	↓ egg production, estrogenic-like activity similar to or greater than BPA	[77]
		96 h (embryo–larva)	1–100 µg/L	↑ ER and aromatase gene activity, altered reproductive neuroendocrine genes	[78]
BPS	Danio rerio	4 hpf–120 dpf	1 and 100 µg/L	female-dominant sex ratio, impaired reproductive capacity	[61]
		75 days	0.1, 1, 10, 100 μg/L	↑ female to male sex ratio, ↑ estradiol and VTG (males, females), ↓ egg production	[79]
	Danio rerio, male	96 hpf–6 dpf	1–200 mg/L	↑ VTG at high concentrations, induced estrogenic response, less potent than BPA	[50,69]
		21 days	8, 40, 200 μg/mL	↑ VTG, ↑ aromatase and estradiol	[52]

Research further indicates that BPs can bind to specific estrogen receptors, particularly estrogen receptor alpha (ERα), thereby either activating or inhibiting estrogenic pathways. For instance, BPA disrupts steroidogenesis in Leydig cells, leading to an imbalance in sex hormones, which is critical for reproductive health [80]. This disruption manifests as altered hormone levels, specifically increased estradiol and decreased testosterone, with profound effects on reproductive function. Furthermore, exposure to BPA has been associated with changes in the expression of genes related to estrogen signalling, such as hoxa10, which is crucial for uterine development and function [81]. Moreover, the estrogenic effects of BPs have been observed in various animal models. For example, exposure to BPS has been linked to reproductive impairments and hormonal imbalances in zebrafish (Danio rerio), highlighting the potential for these compounds to disrupt endocrine functions across species [79].

The assessment of estrogenic disruption can be conducted at multiple biological levels, each providing critical insights into endocrine interference. Gene expression analyses, particularly the measurement of VTG levels, serve as a reliable biomarker for estrogenic activity. VTG is a precursor protein synthesised in response to estrogen, making it a key indicator of endocrine disruption in male fish. Histological examinations of gonadal tissues can reveal alterations in sex differentiation and reproductive structures. Additionally, assessments of egg quality, hatching success, and behavioural studies (e.g., changes in mating and spawning behaviour) help elucidate the broader ecological consequences of estrogenic disruption [82,83,84].

BPA has been extensively studied and is known to induce VTG production in male zebrafish (Danio rerio), resulting in altered reproductive behaviours and developmental anomalies, including reduced hatching success [55,69]. Similarly, BPS exhibits estrogenic activity, leading to delayed hatching and increased malformation rates in embryos. While BPF is generally less potent than BPA, it still demonstrates significant estrogenic effects, impacting reproductive health and egg quality [82]. BPAF, on the other hand, is more potent than BPA, causing severe reproductive and developmental disruptions, including significant effects on egg quality and hatching success [78].

Furthermore, studies indicate that other BPs, such as bisphenol B (BPB) and bisphenol C (BPC), also possess estrogenic activity, although research on their specific effects remains limited [15,85]. The presence of these compounds in aquatic environments poses a significant risk to biodiversity and ecosystem health, reinforcing the need for comprehensive evaluations of their endocrine-disrupting potential.

The estrogenic activity of BPs has been well-documented in numerous studies involving various fish species. Even at low concentrations, exposure to BPs has been shown to induce the production of VTG, a female-specific egg yolk precursor protein, in male fish, a clear indication of feminisation and endocrine disruption. Furthermore, bisphenol exposure has been linked to altered sexual development, including the development of intersex characteristics, reduced spawning behaviour, and impaired reproductive success in both male and female fish [4,21].

Diagnosing estrogenic disruption caused by BPA and its analogues involves a comprehensive approach that integrates biochemical assays, in vivo studies, and molecular analyses. One primary method for assessing estrogenic activity is through cell-based assays that measure the binding affinity to estrogen receptors (ERs). Research has demonstrated that BPA and its analogues can activate both estrogen receptor alpha (ERα) and estrogen receptor beta (ERβ) in vitro, indicating their potential to disrupt normal hormonal signalling [79,86]. Techniques such as luciferase reporter assays are effective for quantifying the transcriptional activation of ERs by these compounds, providing clear evidence of their estrogenic potential [87].

In addition to biochemical assays, animal models—particularly rodents and zebrafish (Danio rerio)—are crucial for understanding the physiological effects of bisphenol exposure. The uterotrophic assay, which measures changes in uterine weight in response to estrogenic compounds, has long been a standard method for evaluating estrogenicity in vivo. Furthermore, studies involving zebrafish embryos exposed to bisphenols have shown alterations in primordial germ cell migration and other developmental anomalies, which can serve as indicators of estrogenic disruption [50]. These in vivo models allow researchers to observe the systemic effects of bisphenols and their impact on reproductive health.

Finally, understanding the molecular mechanisms by which bisphenols exert their estrogenic effects is essential for accurate diagnosis. Research indicates that bisphenols can alter the expression of genes related to estrogen signalling pathways, including the upregulation of VTG in fish models [7]. Moreover, bisphenols may induce epigenetic changes that affect hormone receptor expression and function, contributing to long-term disruptions in endocrine signalling [83,88].

Additionally, the impact of BPs on development and behaviour, particularly in model organisms like zebrafish (Danio rerio), serves as a critical indicator. Research has shown that exposure to BPs can result in alterations in brain development and reproductive behaviours, which are linked to disrupted estrogen signalling [69,89]. These developmental effects may serve as biomarkers for estrogenic disruption. A growing body of research highlights that not only bisphenol BPA but also its analogues—including BPS, BPF, BPAF, and others—exhibit estrogenic properties and can lead to reproductive impairments in fish.

4.2. Androgenic Activity and Disruption

In addition to their estrogenic effects, BPs have also been shown to exhibit androgenic activity, which can disrupt the normal functioning of the androgen system in fish [90]. Androgen disruption in fish can manifest through a range of physical, behavioural, and physiological symptoms. The effects of such disruption can vary among species and depend on the timing and extent of exposure. An overview of significant findings on the androgenic effects of selected BPs is given in Table 5.

Table 5. Overview of significant findings on the androgenic effects of selected bisphenols in fish (↑ = increase; ↓ = decrease; dpf = days post-fertilisation; hpf = hours post-fertilisation).

Bisphenol Type	Fish Species	Exposure Duration	Concentration	Effects Observed	References
BPA	Carassius auratus	7–90 days	0.2 and 20 µg/L	↓ sperm quality (number, motility, volume), ↓ testosterone, antiandrogenic effects	[74]
BPS	Danio rerio	21 days	8, 40, 200 μg/mL	↓ endogenous androgens (males)	[52]
		21 days	0.5, 5, 50 μg/L	antiandrogenic effect, ↓ gonad weight (males and females)	[79]
		75 days	0.1, 1, 10, 100 μg/L	↓ testosterone (males), ↓ gonad weight, ↓ testosterone (males), ↓ sperm count	[79]
		4 hpf–120 dpf	1 and 100 µg/L	altered steroid hormone levels	[61]
	Gadus morhua, cells		EC50 25.0 µmol/L; tested in the range of 0.003–50 µmol/L	↑ activity of androgen receptor gmAR	[15]
BPF	Danio rerio	21 days	0.1 and 1 mg/L	↓ testosterone (males), antiandrogenic effects	[75]
BPAF	Gadus morhua, cells		EC50 25.0 µmol/L; tested in the range of 0.003–50 µmol/L	↑ activation of the androgen receptor gmAR	[15]
	Danio rerio	6 dpf	0.1 and 1 µmol/L	↓ locomotor activity, ↑ aromatase B (estrogenic/antiandrogenic marker)	[91]
		4 hpf–120 dpf	5, 25, 125 μg/L	↓ testosterone (males), impaired parental sperm quality	[92]

Table 5. Overview of significant findings on the androgenic effects of selected bisphenols in fish (↑ = increase; ↓ = decrease; dpf = days post-fertilisation; hpf = hours post-fertilisation).

Bisphenol Type	Fish Species	Exposure Duration	Concentration	Effects Observed	References
BPA	Carassius auratus	7–90 days	0.2 and 20 µg/L	↓ sperm quality (number, motility, volume), ↓ testosterone, antiandrogenic effects	[74]
BPS	Danio rerio	21 days	8, 40, 200 μg/mL	↓ endogenous androgens (males)	[52]
		21 days	0.5, 5, 50 μg/L	antiandrogenic effect, ↓ gonad weight (males and females)	[79]
		75 days	0.1, 1, 10, 100 μg/L	↓ testosterone (males), ↓ gonad weight, ↓ testosterone (males), ↓ sperm count	[79]
		4 hpf–120 dpf	1 and 100 µg/L	altered steroid hormone levels	[61]
	Gadus morhua, cells		EC50 25.0 µmol/L; tested in the range of 0.003–50 µmol/L	↑ activity of androgen receptor gmAR	[15]
BPF	Danio rerio	21 days	0.1 and 1 mg/L	↓ testosterone (males), antiandrogenic effects	[75]
BPAF	Gadus morhua, cells		EC50 25.0 µmol/L; tested in the range of 0.003–50 µmol/L	↑ activation of the androgen receptor gmAR	[15]
	Danio rerio	6 dpf	0.1 and 1 µmol/L	↓ locomotor activity, ↑ aromatase B (estrogenic/antiandrogenic marker)	[91]
		4 hpf–120 dpf	5, 25, 125 μg/L	↓ testosterone (males), impaired parental sperm quality	[92]

The androgenic activity of BPs can disrupt the endocrine system by mimicking or inhibiting natural androgens, leading to alterations in sexual development and reproduction [79,93]. For instance, studies have documented changes in male secondary sexual characteristics, aggression, and reproductive behaviours in fish exposed to BPs [55,94]. The mechanisms by which BPs disrupt androgen receptor (AR) pathways involve complex molecular interactions. These compounds can act as antagonists at the AR, potentially interfering with the receptor’s normal function, as well as altering gene expression and signalling pathways that regulate androgen-mediated processes [95,96].

One of the most notable indicators of androgen disruption in fish is, as said, the alteration in sexual characteristics and gonadal morphology. Male fish may exhibit feminisation, with reduced testis size, the presence of ovarian tissue, or the development of female secondary sexual characteristics [97,98]. For example, studies have shown that exposure to androgens and androgen mimetics can lead to the masculinisation of phenotypes in some species, including increased growth of male-specific anatomical features, while concomitantly reducing ovarian tissue or VTG expression as mentioned in the previous section [97,99,100,101].

Behaviour can also be significantly altered as a result of androgen disruption. For example, males exposed to androgenic EDCs have been reported to demonstrate decreased aggression, altered courtship behaviours, and compromised reproductive interactions. These behavioural changes can be linked to the endocrine changes induced by androgens, which affect brain function and are associated with neurogenesis [98,102]. Changes in social dynamics are particularly evident in species like zebrafish (Danio rerio), where exposure to androgenic substances can affect dominant/submissive interactions among males [98].

At the molecular level, disruptions in gene expression patterns that usually are regulated by androgens may occur. Key genes involved in reproduction, such as those encoding steroidogenic enzymes and receptors, may become upregulated or downregulated [99,103]. For example, research has demonstrated that exposure to androgenic substances can lead to significant changes in the expression of genes associated with spermatogenesis and testosterone synthesis by Leydig cells, critical for proper testicular function [104]. Moreover, alterations in the expression of aromatase, which is responsible for converting androgens to estrogens, may occur, leading to disrupted endocrine balance [101,105]. Gene expression profiling studies reveal that alterations induced by androgenic and antiandrogenic compounds significantly modulate the transcription of numerous genes, with potential downstream physiological implications, such as infertility and reduced fitness, in affected populations [100,106].

These changes could have broader ecosystem-level implications [43,107]. The ecological consequences of these androgenic disruptions can be profound. Altered sex ratios and hindered reproductive capacity in populations may lead to declines in biodiversity and changes in ecosystem dynamics [55,85]. Furthermore, the widespread environmental presence of BPs necessitates robust regulatory frameworks to monitor and limit their usage, addressing the dual challenges posed by their known estrogenic and androgenic activities [9,79].

Detection of androgenic disruption due to BPs can be achieved through various methods. In vitro assays using cultured cells, such as prostate cancer cell lines, allow for the assessment of androgenic and antiandrogenic activities of BPs by measuring changes in cellular proliferation and transcriptomic profiles [108]. In vivo models, particularly zebrafish (Danio rerio) and other fish species, are also utilised to observe phenotypic changes linked to androgen disruption, including alterations in growth patterns and reproductive behaviours [53,94]. Advanced techniques such as molecular docking and in silico modelling provide further insight into the binding affinities of BPs to androgen receptors and can inform the development of substitutes that might mitigate these effects [95,109].

The link between gene expression and androgenic disruption, particularly due to environmental contaminants like BPs, is a complex field that underscores how endocrine disruptors can alter reproductive and developmental pathways through modulation of gene regulation. Androgens, which play a crucial role in various physiological processes, influence gene expression by interacting with the androgen receptor (AR). When BPs like BPA and its analogues disrupt this signalling, they can lead to significant changes in gene expression that may alter developmental and reproductive functions across species [85,93]. Androgens exert their effects primarily through the androgen receptor, a transcription factor that, upon binding to its ligand, undergoes a conformational change, allowing it to translocate into the nucleus and bind to androgen response elements (AREs) in the promoters of target genes [110]. This binding promotes the transcription of several genes involved in crucial biological processes, including cell proliferation, differentiation, and survival [110,111]. In healthy conditions, this regulatory mechanism ensures that physiological levels of androgens stimulate the appropriate expression of genes necessary for male sexual differentiation and reproductive function [112]. However, exposure to BPs can disrupt this normal function. These compounds can act as antiandrogens, blocking the action of androgens on the receptor and thereby modulating the expression of genes that depend on AR activity. For instance, studies have shown that bisphenol exposure can lead to the repression of several genes critical for male reproductive health, as the androgen receptor’s ability to activate gene transcription is compromised by these environmental pollutants [9,94]. BPs have been implicated in altering gene expression profiles related to steroidogenesis and spermatogenesis, suggesting that the disruption can have direct repercussions on male fertility and reproductive behaviours [93,109,113]. In addition to these direct effects, androgen signalling can also intersect with other signalling pathways, such as the Wnt/β-catenin pathway, to influence gene expression. For instance, certain genes involved in these pathways are downregulated in response to hormonal disruptions caused by EDCs, further complicating the resultant gene expression landscape [114]. Disruption in AR signalling leads not only to decreased expression of androgen-responsive genes but also to the upregulation of stress response genes, which may act as compensatory mechanisms [1,114]. Such alterations can contribute to developmental anomalies and reproductive diseases via feedback mechanisms that engage various signalling cascades affected by initial endocrine disruption [85,94,113].

4.3. Thyroid Activity and Disruption

Thyroid disruption refers to the interference of normal thyroid gland function by chemical substances that affect the production, release, transport, or action of thyroid hormones in an organism. Studies have shown that BP exposure is associated with thyroid hormone levels and can induce thyroid autoimmunity, leading to thyroid dysfunction. Thyroid hormones play a critical role in regulating a wide range of physiological processes, including growth, development, and metabolism. Disruption of thyroid hormone homeostasis by BPs can lead to impaired growth, delayed development, and other physiological abnormalities in exposed fish [43]. An overview of significant findings on the thyroid effects of selected BPs is given in Table 6 and Table 7. The thyroid-disrupting potential of BPs is linked to several molecular mechanisms. BPs are known to act as antagonists of thyroid hormone receptors (TRs), interfering with the normal signalling pathways regulated by these hormones [13]. BPA, in particular, has been shown to bind to TRs, especially the TRβ isoform, potentially leading to altered gene expression and downstream biological effects [12,115,116]. Additionally, BPs can inhibit the synthesis of thyroid hormones by impairing iodine uptake and reducing the activity of thyroid peroxidase (TPO), an enzyme essential for hormone production [12,117]. These disruptions affect the thyroid axis as hormone synthesis depends on the transport of iodide into thyrocytes [47].

Experimental studies have further supported these findings, showing that BPs can suppress the stimulatory effects of thyroid hormones in a dose-dependent manner. Research has demonstrated that BPs disrupt the expression of genes involved in thyroid function, including thyrotropin (TSH) and deiodinases [118,119]. Analysis of BPA metabolites revealed that, while the 3-hydroxyl metabolite exhibits estrogenic activity, the glucuronide metabolite does not, indicating that bioactivities vary depending on the metabolic transformation [120].

Table 6. Overview of significant findings on the thyroid effects of selected bisphenols in fish (↑ = increase; ↓ = decrease; 3,5-T2 = 3,5-diiodothyronine; dpf = days post-fertilisation; hpf = hours post-fertilisation; HPG = hypothalamic-pituitary–gonadal axis; HPT = Hypothalamic-pituitary-thyroid axis; T3 = triiodothyronine; T4 = thyroxine; TTR = Transthyretin).

Bisphenol Type	Fish Species	Exposure Duration	Concentration	Effects Observed	References
BPA	Danio rerio	120 hpf	0.4, 2, 10 mg/L	↑ T3/T4, altered thyroid development/transport/metabolism genes, delayed hatching	[56]
		Up to 8 dpf	0.2, 0.6, 1.3, 2.8 mg/L	↓ T4, ↓ 3,5-T2, disrupted THS homeostasis, altered retinal morphology	[121]
		Up to 8 dpf	0.25, 0.5, 1.2, 4 µg/L	TRβ antagonism, inhibition of deiodinases, and altered phase II enzyme transcripts	[122]
		96 hpf	1 and 100 µg/L	↓ heart rate, ↑ SV-BA distance, altered dio3b, thrβ, myh7 gene expression	[123]
		96 hpf–6 dpf	1–200 mg/L	developmental deformities	[50,69]
	Oryzias melastigma	70 days	200 µg/L	↓ hatching rate, altered HPG axis gene expression, epigenetic changes in offspring	[73]
	Pimephales promelas	4 days	16–1280 µg/L	↓ hatchability in F1 generation	[69]
BPS	Danio rerio	7 days (adult)	100 µg/L	↑ TTR protein (plasma, liver, brain), ↑ T3/T4, thyroid tissue damage, altered HPT axis gene expression	[124]
		75 days	0.1, 1, 10, 100 μg/L	↓ body length and weight, ↓ T3 and T4, ↑ liver weight, ↓ hatching rate, ↑ time to hatch	[79]
		96 h	1 and 100 µg/L	Promoted heart pumping, altered thrβ, myh7 gene expression	[123]
		4 hpf–120 dpf	1 and 100 µg/L	delayed hatching, ↓ offspring survival	[61]

Table 6. Overview of significant findings on the thyroid effects of selected bisphenols in fish (↑ = increase; ↓ = decrease; 3,5-T2 = 3,5-diiodothyronine; dpf = days post-fertilisation; hpf = hours post-fertilisation; HPG = hypothalamic-pituitary–gonadal axis; HPT = Hypothalamic-pituitary-thyroid axis; T3 = triiodothyronine; T4 = thyroxine; TTR = Transthyretin).

Bisphenol Type	Fish Species	Exposure Duration	Concentration	Effects Observed	References
BPA	Danio rerio	120 hpf	0.4, 2, 10 mg/L	↑ T3/T4, altered thyroid development/transport/metabolism genes, delayed hatching	[56]
		Up to 8 dpf	0.2, 0.6, 1.3, 2.8 mg/L	↓ T4, ↓ 3,5-T2, disrupted THS homeostasis, altered retinal morphology	[121]
		Up to 8 dpf	0.25, 0.5, 1.2, 4 µg/L	TRβ antagonism, inhibition of deiodinases, and altered phase II enzyme transcripts	[122]
		96 hpf	1 and 100 µg/L	↓ heart rate, ↑ SV-BA distance, altered dio3b, thrβ, myh7 gene expression	[123]
		96 hpf–6 dpf	1–200 mg/L	developmental deformities	[50,69]
	Oryzias melastigma	70 days	200 µg/L	↓ hatching rate, altered HPG axis gene expression, epigenetic changes in offspring	[73]
	Pimephales promelas	4 days	16–1280 µg/L	↓ hatchability in F1 generation	[69]
BPS	Danio rerio	7 days (adult)	100 µg/L	↑ TTR protein (plasma, liver, brain), ↑ T3/T4, thyroid tissue damage, altered HPT axis gene expression	[124]
		75 days	0.1, 1, 10, 100 μg/L	↓ body length and weight, ↓ T3 and T4, ↑ liver weight, ↓ hatching rate, ↑ time to hatch	[79]
		96 h	1 and 100 µg/L	Promoted heart pumping, altered thrβ, myh7 gene expression	[123]
		4 hpf–120 dpf	1 and 100 µg/L	delayed hatching, ↓ offspring survival	[61]

Table 7. Overview of significant findings on the thyroid effects of selected bisphenols in fish (↑ = increase; ↓ = decrease; 3,5-T2 = 3,5-diiodothyronine; dpf = days post-fertilisation; hpf = hours post-fertilisation; HPG = hypothalamic-pituitary–gonadal axis; HPT = Hypothalamic-pituitary–thyroid axis; T3 = triiodothyronine; T4 = thyroxine).

Bisphenol Type	Fish Species	Exposure Duration	Concentration	Effects Observed	References
BPF	Danio rerio	7 days (adult)	10 and 100 µg/L	↑ TTR protein (plasma, liver, brain), ↑ T3/T4, thyroid follicle pathology, altered HPT axis gene expression	[124]
		120 hpf (larvae)	2 mg/L	↑ T3/T4, altered thyroid development/transport/metabolism genes, delayed hatching	[56]
		96 h	1 and 100 µg/L	promoted heart pumping, altered thrβ, myh7 gene expression	[123]
		96 hpf–6 dpf	1–200 mg/L	developmental deformities	[50,69]
		21 days	0.1 and 1 mg/L	disrupted HPG axis gene expression	[75]
	Oryzias melastigma	70 days	200 µg/L	↓ hatching rate, altered HPG axis gene expression, epigenetic changes in offspring	[73]
BPAF	Danio rerio	4 hpf–120 dpf	5, 25, 125 µg/L	↑ malformations and lower survival in offspring, delayed hatching, altered HPG axis and liver gene expression	[61]
		96 hpf–6 dpf	1–200 mg/L	developmental deformities	[50,69]
		4 hpf–120 dpf	5, 25, 125 μg/L	↑ malformations, ↓ survival rate in offspring, ↓ hatching at 5 μg/L BPAF	[92]
	Oryzias melastigma	70 days	200 µg/L	↓ hatching rates, altered HPG axis gene expression, epigenetic changes in offspring	[73]
BPB	Danio rerio	21 days	1 mg/L	↓ hatching rate and viability	[77]
BPZ	Danio rerio	120 hpf	0.18 and 2.9 µg/L	disrupted T3/T4, altered thyroid-related gene expression, delayed hatching	[56]
TBBPA	Danio rerio	5 dpf	100, 200, 300, 400 µg/L	inhibition of deiodinases, altered gene expression	[125,126]

Table 7. Overview of significant findings on the thyroid effects of selected bisphenols in fish (↑ = increase; ↓ = decrease; 3,5-T2 = 3,5-diiodothyronine; dpf = days post-fertilisation; hpf = hours post-fertilisation; HPG = hypothalamic-pituitary–gonadal axis; HPT = Hypothalamic-pituitary–thyroid axis; T3 = triiodothyronine; T4 = thyroxine).

Bisphenol Type	Fish Species	Exposure Duration	Concentration	Effects Observed	References
BPF	Danio rerio	7 days (adult)	10 and 100 µg/L	↑ TTR protein (plasma, liver, brain), ↑ T3/T4, thyroid follicle pathology, altered HPT axis gene expression	[124]
		120 hpf (larvae)	2 mg/L	↑ T3/T4, altered thyroid development/transport/metabolism genes, delayed hatching	[56]
		96 h	1 and 100 µg/L	promoted heart pumping, altered thrβ, myh7 gene expression	[123]
		96 hpf–6 dpf	1–200 mg/L	developmental deformities	[50,69]
		21 days	0.1 and 1 mg/L	disrupted HPG axis gene expression	[75]
	Oryzias melastigma	70 days	200 µg/L	↓ hatching rate, altered HPG axis gene expression, epigenetic changes in offspring	[73]
BPAF	Danio rerio	4 hpf–120 dpf	5, 25, 125 µg/L	↑ malformations and lower survival in offspring, delayed hatching, altered HPG axis and liver gene expression	[61]
		96 hpf–6 dpf	1–200 mg/L	developmental deformities	[50,69]
		4 hpf–120 dpf	5, 25, 125 μg/L	↑ malformations, ↓ survival rate in offspring, ↓ hatching at 5 μg/L BPAF	[92]
	Oryzias melastigma	70 days	200 µg/L	↓ hatching rates, altered HPG axis gene expression, epigenetic changes in offspring	[73]
BPB	Danio rerio	21 days	1 mg/L	↓ hatching rate and viability	[77]
BPZ	Danio rerio	120 hpf	0.18 and 2.9 µg/L	disrupted T3/T4, altered thyroid-related gene expression, delayed hatching	[56]
TBBPA	Danio rerio	5 dpf	100, 200, 300, 400 µg/L	inhibition of deiodinases, altered gene expression	[125,126]

The impact of BPs on thyroid function is particularly concerning during critical periods of growth and development. Human studies have documented that exposure to BPs during sensitive windows, such as gestation and early childhood, can result in lasting disruptions in thyroid function, which may affect both cognitive and physical development [127,128]. In aquatic models, research on zebrafish (Danio rerio) has shown that waterborne exposure to BPS disrupts thyroid hormone signalling pathways, leading to developmental abnormalities [54]. Furthermore, population-based studies have associated exposure to endocrine-disrupting chemicals (EDCs), such as BPA, with thyroid dysfunction, underlining the need for regulatory scrutiny [129,130].

Finally, BPA exposure has been shown to have variable effects on thyroid hormone levels in both human and animal studies, including increased T4 levels and alterations in thyroid gland weight and histology. These alterations are consistent with its ability to interfere with gene expression related to hormone synthesis and receptor-mediated action [66,90]. It was shown that in zebrafish (Danio rerio) BPA, BPF, and BPS can bind TRβ and exert antagonistic activity. For example, BPF exposure altered T₄, T₃, and TSH levels and changed the expression of genes including Tg, Ttr, and Ugt1ab. BPS exposure decreased T₄ and T₃ levels and increased TSH levels. Furthermore, in zebrafish, BPS treatment increased the expression of genes, including Ttr and Ugt1ab [12].

Thyroid disruption in fish, particularly in model species such as zebrafish (Danio rerio), can be assessed through various methodological approaches. Detecting alterations in thyroid function is essential for understanding the ecological impact of these contaminants and for protecting aquatic organisms. One of the most direct approaches involves measuring thyroid hormone levels, such as triiodothyronine (T3) and thyroxine (T4), in plasma or tissue samples from exposed fish. Alterations in these hormone concentrations serve as indicators of disrupted thyroid function. For example, Liu et al. (2022) reported that long-term exposure to TDCPP significantly altered T4 levels in zebrafish (Danio rerio) [131].

Gene expression analysis also plays a critical role in detecting thyroid disruption. Techniques such as quantitative PCR (qPCR) and RNA sequencing are used to assess the expression of genes involved in thyroid hormone regulation, including dio1 and dio2, which are responsible for the conversion of T4 to its more active form, T3. Studies have demonstrated that exposure to various substances alters the expression of thyroid receptor genes in zebrafish larvae, confirming the occurrence of thyroid disruption [125].

Histological examination of thyroid follicles is another important tool. Structural changes in the thyroid gland can reveal damage induced by EDCs. Kraft et al. (2022) showed that exposure to triclosan and benzophenone-2 resulted in morphometric changes in zebrafish (Danio rerio) thyroid follicles, highlighting them as an effective indicator of thyroid disruption [132].

In addition to histology, researchers often monitor developmental abnormalities such as underdeveloped eyes or swim bladders, which may signal disrupted thyroid hormone function. These phenotypic outcomes are particularly relevant during early developmental stages when thyroid hormones play a key regulatory role. While specific references are not always cited for these broader findings, they are supported by studies like Baumann et al. (2016) [125].

In vivo testing using zebrafish embryos and larvae is widely employed to study the effects of EDCs. For instance, Tang et al. (2015) demonstrated that developmental exposure to BPAF disrupted the hypothalamic–pituitary–thyroid (HPT) axis, offering direct evidence of thyroid disruption [133].

Transgenic zebrafish (Danio rerio) models further enhance detection sensitivity. Some lines are engineered with thyroid response elements that trigger the expression of green fluorescent protein (GFP) in response to thyroid hormone activity. This visual and quantifiable response enables effective identification of disrupted thyroid signalling [134].

The development and application of adverse outcome pathway (AOP) frameworks provide a structured means to link molecular or cellular changes caused by EDCs to adverse biological outcomes. AOPs help identify key events and sensitive endpoints, thereby enhancing the ability to assess thyroid disruption systematically. Studies illustrated how AOP networks can be used to correlate environmental exposure with specific endocrine outcomes, offering valuable guidance for regulatory decision making [135,136].

4.4. Glucocorticoid Activity and Disruption

BPs have attracted increasing attention due to their potential to interfere with glucocorticoid hormones, most notably cortisol. Cortisol plays a central role in stress responses, energy metabolism, immune regulation, and other vital physiological processes in vertebrates, including fish [9,137]. In aquatic organisms, exposure to BPs has been linked to altered cortisol levels and disrupted enzyme activity related to glucocorticoid metabolism. These effects can impair the function of the hypothalamic–pituitary–adrenal (HPA) or hypothalamic–pituitary–interrenal (HPI) axis, ultimately compromising the fish’s ability to regulate stress [138,139]. Significant findings on the glucocorticoid disruption changes of BPs are shown in Table 8.

One primary mechanism of endocrine disruption involves the binding of BPs to glucocorticoid receptors (GRs). BPA, in particular, can mimic the action of endogenous glucocorticoids by binding to GRs, thereby disrupting hormonal homeostasis and impairing normal stress regulation. Prolonged or excessive exposure to BPA has been associated with physiological imbalances, including altered cortisol and glucose levels, which can negatively affect metabolic regulation and overall fish health and fitness [3,120].

BPs exert their disruptive effects primarily by interfering with hormone signalling pathways. Through their interaction with nuclear hormone receptors such as GRs, BPs can inhibit or modify the normal transcriptional activity driven by natural glucocorticoids. This antagonistic or agonistic activity alters gene expression and can compromise stress responses and immune function [129].

In addition to receptor-level interactions, BPs can interfere with the biosynthesis of glucocorticoids. These chemicals regulate the expression and function of genes and enzymes involved in the production of steroid hormones. EDCs, including BPs, have been shown to inhibit key enzymes required for cortisol synthesis, leading to dysregulated hormone levels and impaired stress responses [117,119].

BPs also affect the regulation of the HPI axis in fish, disrupting hormonal feedback mechanisms and altering levels of corticotropin-releasing hormone (CRH) and adrenocorticotropic hormone (ACTH). These hormones are essential for proper cortisol regulation, and their dysregulation may contribute to chronic stress conditions and impaired physiological homeostasis [119].

Table 8. Overview of significant findings on the glucocorticoid disruption of selected bisphenols in fish (↑ = increase; ↓ = decrease; hpf = hours post-fertilisation; HPI = hypothalamic–pituitary–interrenal).

Bisphenol Type	Fish Species	Exposure Duration	Concentration	Effects Observed	References
BPA	Danio rerio	96 h	1500 µg/L	↑ anxiety-like behaviour, neurotoxic effects	[140]
		30 days	1, 10, 100 µg/L	↑ oxidative stress	[3]
	Danio rerio, male	6 weeks	100 and 2000 µg/L	↑ fat deposition, ↑ body weight, ↑ lipid synthesis, inflammation, antioxidant response	[70]
	Cyprinus carpio, juvenile	30 days	0.1, 1, 10, 100, 1000 µg/L	(≥0.1 μg/L) ↓ immune response, ↑ oxidative stress	[141]
	Ctenopharyngodon idella, ovary cells	48 h (in vitro)	30 μmol/L	↑ oxidative stress, altered DNA methylation	[142]
	Aristichthys nobilis	60 days	0.1, 1, 10 µg/L	↑ oxidative stress, behavioural changes, and physiological disturbances	[58]
BPS	Danio rerio	120 days (embryo to adult)	1, 10, 100 µg/L	↑ whole-body cortisol, altered HPI axis gene expression, ↑ anxiety-like behaviour	[62]
		75 days	1, 10, 30 µg/L	impaired anxiety/fear responses, altered antioxidant gene expression	[143]
	Danio rerio, adult female	120 days	1, 10, 30 µg/L	1 μg/L improved cognitive behaviours, 10/30 μg/L impaired cognitive behaviours, altered glutamatergic signalling	[144]
	Cyprinus carpio, juvenile	60 days	1, 10, 100 µg/L	↑ oxidative stress, chronic inflammatory stress in the liver	[141]
BPAF	Danio rerio, embryo/adult	120 hpf	5, 50, 500 µg/L	altered anxiety-like/aggressive behaviour, ↑ oxidative stress	[92]

Table 8. Overview of significant findings on the glucocorticoid disruption of selected bisphenols in fish (↑ = increase; ↓ = decrease; hpf = hours post-fertilisation; HPI = hypothalamic–pituitary–interrenal).

Bisphenol Type	Fish Species	Exposure Duration	Concentration	Effects Observed	References
BPA	Danio rerio	96 h	1500 µg/L	↑ anxiety-like behaviour, neurotoxic effects	[140]
		30 days	1, 10, 100 µg/L	↑ oxidative stress	[3]
	Danio rerio, male	6 weeks	100 and 2000 µg/L	↑ fat deposition, ↑ body weight, ↑ lipid synthesis, inflammation, antioxidant response	[70]
	Cyprinus carpio, juvenile	30 days	0.1, 1, 10, 100, 1000 µg/L	(≥0.1 μg/L) ↓ immune response, ↑ oxidative stress	[141]
	Ctenopharyngodon idella, ovary cells	48 h (in vitro)	30 μmol/L	↑ oxidative stress, altered DNA methylation	[142]
	Aristichthys nobilis	60 days	0.1, 1, 10 µg/L	↑ oxidative stress, behavioural changes, and physiological disturbances	[58]
BPS	Danio rerio	120 days (embryo to adult)	1, 10, 100 µg/L	↑ whole-body cortisol, altered HPI axis gene expression, ↑ anxiety-like behaviour	[62]
		75 days	1, 10, 30 µg/L	impaired anxiety/fear responses, altered antioxidant gene expression	[143]
	Danio rerio, adult female	120 days	1, 10, 30 µg/L	1 μg/L improved cognitive behaviours, 10/30 μg/L impaired cognitive behaviours, altered glutamatergic signalling	[144]
	Cyprinus carpio, juvenile	60 days	1, 10, 100 µg/L	↑ oxidative stress, chronic inflammatory stress in the liver	[141]
BPAF	Danio rerio, embryo/adult	120 hpf	5, 50, 500 µg/L	altered anxiety-like/aggressive behaviour, ↑ oxidative stress	[92]

Changes in glucocorticoid signalling caused by bisphenol exposure can directly affect glucose metabolism. Glucocorticoids are crucial regulators of gluconeogenesis and glucose mobilisation. Chronic or excessive glucocorticoid activity, particularly when triggered by EDCs like BPA, can lead to metabolic disturbances such as insulin resistance and disrupted energy balance. These effects are especially harmful during critical periods of development [145].

Experimental studies support the hypothesis that BPs act as GR agonists. For instance, Kouche et al. (2025) demonstrated that BPA significantly activated GRs in cell lines lacking endogenous receptors. The study observed increased transcription of GR-regulated genes, highlighting the capacity of BPA to modulate gene expression pathways related to metabolism and stress responses [146]. Macíková et al. (2014) similarly documented changes in corticosteroid signalling pathways in fish exposed to bisphenols, indicating disruptions to cortisol production and downstream physiological effects, such as altered osmoregulation and immune responses [147].

The disruption of glucocorticoid signalling by bisphenols can produce a range of behavioural and physiological symptoms. Behaviourally, affected fish may show signs of heightened anxiety, increased aggression, reduced foraging, or diminished social interactions—symptoms indicative of impaired stress regulation [126,148]. On a physiological level, bisphenol-induced disruption of glucocorticoid function can lead to immunosuppression, stunted growth, and increased vulnerability to pathogens. Chronic exposure to BPs can result in prolonged activation of the stress axis, reproductive impairments, and developmental abnormalities [115,149]. Endocrine disruption may also manifest through altered gonadal development and reduced fecundity, as glucocorticoid balance is closely linked to the regulation of reproductive hormones. For instance, Wu et al. (2016) found that zebrafish (Danio rerio) exposed to BPA displayed shifts in reproductive hormone levels, suggesting compromised reproductive fitness [117].

The consequences of glucocorticoid disruption extend beyond individual organisms. Disrupted stress and endocrine function can increase mortality, decrease reproductive output, and alter population dynamics. Given the ecological importance of fish as predators, prey, and contributors to nutrient cycling, such disruptions may cascade through aquatic food webs, compromising biodiversity and ecosystem resilience [129,150].

Detecting glucocorticoid alterations in fish, particularly through the measurement of cortisol levels, plays a crucial role in assessing their physiological stress responses. Cortisol is a primary biomarker of stress, and several methods have been developed to measure its concentration in fish. These methods vary in invasiveness, sensitivity, and suitability for different experimental setups or life stages of the fish. Plasma cortisol measurement is one of the most common techniques, providing direct insight into circulating levels of cortisol in the blood. It typically involves blood sampling, followed by analytical techniques like enzyme-linked immunosorbent assay (ELISA) or radioimmunoassay (RIA) for quantification. Although this method is reliable, it is invasive and can induce stress in the fish itself, which may interfere with the results. Despite this limitation, plasma sampling remains a gold standard for establishing baseline cortisol levels. Research by Xiao et al. (2021) highlighted the significance of plasma cortisol as a reliable physiological index of stress, demonstrating its association with alterations in fish metabolism and behaviour [151].

A non-invasive alternative to blood sampling is faecal cortisol analysis. This method involves collecting faecal samples, which can be analysed using techniques such as liquid chromatography coupled with tandem mass spectrometry (LC-MS/MS). It allows researchers to monitor chronic stress levels without directly handling the fish. This approach has been validated as effective, as it preserves the welfare of the animals while providing valuable insights into stress responses over time. Meling et al. (2022) demonstrated the efficacy of this method in assessing stress in fish after their transfer to seawater, proving its usefulness in both field and laboratory settings [152].

Cortisol levels can also be measured in fish skin mucus. This method has gained attention due to the close correlation between cortisol in plasma and mucus. Since cortisol diffuses from the bloodstream into the mucus, it can be sampled non-invasively to assess stress, making it a less invasive alternative to blood sampling. The method has been demonstrated to accurately reflect cortisol levels, providing a reliable means for stress monitoring without harming the fish. According to Carbajal et al. (2019), the correlation between plasma and mucus cortisol was strong, supporting the method’s applicability for stress evaluations in both controlled and natural environments [153].

Another approach involves whole-body cortisol quantification. In this method, the entire body of the fish is homogenised, and cortisol is extracted and analysed using techniques such as ELISA. This method is particularly useful for smaller fish species or larvae, providing a comprehensive measurement of an individual’s stress response. However, it is destructive, making it unsuitable for studies that require repeated sampling of the same individual. Yeh et al. (2013) applied this method to larval zebrafish, demonstrating its potential for assessing stress in early life stages [154].

Measuring cortisol in the water surrounding fish is a non-invasive technique that provides insight into their stress levels. As cortisol is released into the surrounding environment, its concentration in water can serve as an indicator of stress without direct interaction with the fish. Waterborne cortisol levels can be measured using advanced techniques, such as chromatography. While this method offers a less intrusive way of monitoring stress, factors such as water volume, flow rate, and fish density can influence the results. Cao et al. (2017) showed that cortisol concentrations in water correlated with fish biomass and stress levels, making this method particularly useful for monitoring fish welfare in aquaculture systems [155].

Advances in sensor technology have led to the development of carbon nanotube-based biosensors that enable real-time, highly sensitive detection of plasma cortisol levels in fish. These sensors represent a promising innovation for rapid cortisol measurement, offering the advantage of being minimally invasive and providing researchers with efficient monitoring tools. While these sensors are still in the early stages of development for widespread field use, they hold significant potential for enhancing stress monitoring in aquatic environments. Wu et al. (2015) demonstrated the utility of carbon nanotube-enhanced immunosensors for cortisol detection, illustrating their potential for high-throughput and real-time stress monitoring [156].

Available sources indicate that bisphenols (such as BPA, BPS, and BPAF) can independently, but often concurrently, induce both oxidative stress and disruption of glucocorticoid systems [46,60,143]. Oxidative stress refers to a state in which there is an imbalance in an organism between the production of reactive oxygen species (ROS)—unstable oxygen-containing molecules—and the organism’s ability to neutralise them using antioxidants [46,142]. When ROS production exceeds antioxidant defences, cellular and tissue damage may occur. Reactive oxygen species include, among others, superoxide anion (O₂⁻), hydroxyl radical (OH⁻), hydrogen peroxide (H₂O₂), and singlet oxygen (O₂) [142].

Although the sources do not explicitly state a direct causal link between glucocorticoid disruption and oxidative stress (or vice versa) as the primary mechanism of bisphenol toxicity, it is essential to note that both are significant adverse effects of bisphenol exposure. These contribute to overall toxicity and disruption of the endocrine system within an organism. Glucocorticoids play a crucial role in regulating metabolism and immune function [154]. Disruption of these systems can affect overall homeostasis and the organism’s resilience to stress. Oxidative stress, in turn, directly impacts cellular physiology and can lead to cellular and tissue damage [142].

In the broader context of body physiology, stress responses (mediated by glucocorticoids) and oxidative status are interconnected. Stress can trigger ROS production, and oxidative stress can, in turn, influence hormonal signalling and exacerbate the stress response. However, the specific detailed mechanisms of this mutual interaction in the context of bisphenol action are not explicitly described as a primary effect in the provided excerpts.

5. Conclusions

The widespread presence of BPs, particularly in plastics and consumer products, represents a growing concern for both environmental and public health. These compounds, while useful for their versatility in plastics and other industrial applications, have proven to be potent endocrine disruptors that can cause adverse effects on multiple physiological systems. These compounds, especially BPA and its structural analogues, have been shown to act as potent endocrine disruptors, interfering with hormonal systems across multiple axes—including estrogenic, androgenic, thyroidal, and glucocorticoid pathways. In fish, such disruptions lead to reproductive impairments, developmental anomalies, behavioural alterations, and physiological dysfunctions, posing serious threats to individual fitness and population stability. Despite increasing awareness of the risks posed by BPA, its widespread substitution with structurally similar analogues such as BPS, BPF, and BPAF has introduced additional complexity. Emerging evidence suggests these alternatives may exhibit equal or even greater endocrine-disrupting potency, highlighting the inadequacy of simple chemical substitution without thorough toxicological evaluation.

In conclusion, the presence of BPs in our modern world represents a significant health and environmental concern, requiring a comprehensive and multifaceted response. Their widespread use and complex toxicological profiles make effective regulation challenging, necessitating ongoing research, collaboration, and public education. The transition to alternative compounds must consider the potential for similar or even more potent endocrine-disrupting properties, and risk assessment frameworks should account for the potential synergistic effects of multiple disruptors.

It is vital to emphasise the intricacies involved in understanding BPs’ behaviour and distribution in the environment. Understanding the environmental fate, bioaccumulation, and biomagnification of BPs in aquatic ecosystems is essential for assessing long-term ecological risks. Their persistence in sediments, water, and biota enables continuous exposure of fish and other organisms, exacerbating endocrine disruption across trophic levels. Regulatory measures must consider these factors to minimise exposure of fish and other aquatic organisms. Finally, even after regulatory measures have been implemented, ongoing monitoring and evaluation of BPs and their substitutes are crucial to ensure their efficacy and adapt to new information.

While the transition to alternative compounds has begun, this process is fraught with difficulties, as these substitutes may possess similar or even more potent endocrine-disrupting properties. This highlights the importance of ongoing research and interdisciplinary collaboration in fully understanding the toxicological profiles of these compounds and informing effective risk management strategies. Overall, the regulatory landscape for BPs is dynamic and evolving, necessitating a continued focus on developing robust policies that safeguard public health and environmental sustainability.

In addressing the health and environmental concerns posed by BPs, it is crucial to take a multifaceted approach that combines risk assessment and regulation with public awareness and education. Effective communication of the potential risks associated with these compounds, as well as practical steps individuals can take to limit their exposure, is crucial in fostering a more informed and proactive society.

Given the potential for BPs to leach into aquatic environments, leading to adverse effects on fish and other aquatic organisms, regulating these compounds also requires a deeper understanding of their environmental fate and persistence. Strategies that minimise exposure to marine life and mitigate environmental impacts must be prioritised alongside the protection of human health.