Next Article in Journal
Probiotics in Fish Nutrition—Long-Standing Household Remedy or Native Nutraceuticals?
Next Article in Special Issue
Contributions on Lindane Degradation by Microcystis aeruginosa PCC 7806
Previous Article in Journal
Effects of Prometryn Exposure Scenarios on Microcystis aeruginosa Growth and Water Qualities in Incubator Experiments
Previous Article in Special Issue
Microplastics in the Aquatic Environment: Occurrence, Persistence, Analysis, and Human Exposure
Order Article Reprints
Font Type:
Arial Georgia Verdana
Font Size:
Aa Aa Aa
Line Spacing:
Column Width:

Endocrine-Disrupting Compounds: An Overview on Their Occurrence in the Aquatic Environment and Human Exposure

Department of Medicine Surgery and Dentistry, Scuola Medica Salernitana, University of Salerno, Via S. Allende, 84081 Baronissi, Italy
Department of Chemistry and Biology, University of Salerno, Via Giovanni Paolo II, 132-84084 Fisciano, Italy
ISPRA, Italian Institute for Environmental Protection and Research, Via Vitaliano Brancati 60, 00144 Rome, Italy
Andrology Unit and Service of Lifestyle Medicine in Uro-Andrology, Local Health Authority (ASL) Salerno, Coordination Unit of the Network for Environmental and Reproductive Health (EcoFoodFertility Project), Italy “Oliveto Citra Hospital”, Via M. Clemente, 84020 Oliveto Citra, Italy
Department of Biology, University of Rome, Tor Vergata, Via della Ricerca Scientifica 1, 00133 Rome, Italy
Authors to whom correspondence should be addressed.
These authors contributed equally to this paper as first authors.
Water 2021, 13(10), 1347;
Received: 23 February 2021 / Revised: 16 April 2021 / Accepted: 11 May 2021 / Published: 12 May 2021


Endocrine-disrupting compounds (EDCs) as emerging contaminants have accumulated in the aquatic environment at concentration levels that have been determined to be significant to humans and animals. Several compounds belong to this family, from natural substances (hormones such as estrone, 17β-estradiol, and estriol) to synthetic chemicals, especially pesticides, pharmaceuticals, and plastic-derived compounds (phthalates, bisphenol A). In this review, we discuss recent works regarding EDC occurrence in the aquatic compartment, strengths and limitations of current analytical methods used for their detection, treatment technologies for their removal from water, and the health issues that they can trigger in humans. Nowadays, many EDCs have been identified in significant amounts in different water matrices including drinking water, thus increasing the possibility of entering the food chain. Several studies correlate human exposure to high concentrations of EDCs with serious effects such as infertility, thyroid dysfunction, early puberty, endometriosis, diabetes, and obesity. Although our intention is not to explain all disorders related to EDCs exposure, this review aims to guide future research towards a deeper knowledge of EDCs’ contamination and accumulation in water, highlighting their toxicity and exposure risks to humans.

Graphical Abstract

1. Introduction

Human activities have introduced a large number of contaminants of emerging concern (CECs) into the environment on a global scale. This category refers to any chemical discovered in the water cycle that had not previously been detected, and so is not yet regulated by an agency, and often presents at very low concentration levels [1,2,3,4,5,6,7]. CECs include a wide class of different types of organic and inorganic chemical compounds such as disinfection byproducts [8], endocrine disruptors [9,10,11], industrial chemicals, natural toxins, persistent organic pollutants (POPs), brominated flame retardants (BFRs), lifestyle compounds such as caffeine, and artificial sweeteners, pesticides, pharmaceuticals and personal care products (PPCPs), which have the potential to harm biota and humans [12,13,14,15].
Among them, endocrine-disrupting compounds (EDCs), also called endocrine-disrupting chemicals, or simply endocrine disruptors, are xenobiotics compounds mainly present in manufactured products such as children’s toys, plastic bottles, polyvinylchloride pipes, detergents, toothpaste, and cosmetics [16,17]. These chemicals can bind to the body’s endocrine receptors to activate, block, or alter natural hormone synthesis and degradation which occur through a plethora of mechanisms resulting in a “false” lack or abnormal hormonal signals that can increase or inhibit normal endocrine function [18,19]. The term endocrine disruptor was first introduced by Colborn in 1991 [20] and subsequently, the International Programme on Chemical Safety (IPCS) of United Nations Environment Programme (UNEP) and World Health Organization (WHO) in 2002 and 2012 [21], defined it as “an exogenous substance or mixture that alters function(s) of the endocrine system and consequently causes adverse health effects in an intact organism, or its progeny, or (sub) populations” and a potential chemical endocrine disruptor as “an exogenous substance or mixture that possesses properties that might be expected to lead to endocrine disruption in an intact organism, or its progeny, or (sub) populations” [22]. During the last years, there have been many modifications to this definition, but we chose to use the original for this review. In fact, this definition is the commonly accepted one for a substance with endocrine-disrupting properties both by the scientific community and by regulatory bodies around the world [23,24,25]. Based on this approach, a chemical must have a demonstrated adverse effect related by a proof of causality to an endocrine disruption mode-of-action to be defined as an endocrine disruptor. As a consequence, in screening and testing chemicals for the endocrine activity or disruption, it is important to use concentrations that maximize the possibility to find a true endocrine effect, and at the same time avoid doses that cause generalized stress endocrine responses or indirect toxicities [26].
The group of molecules identified as endocrine disruptors is highly heterogeneous and can be divided into synthetic chemicals (from anthropogenic activities e.g., 17α-ethinylestradiol (EE2), pesticides e.g., atrazine, phthalates, alkylphenol ethoxylate surfactants, nonylphenol (NP), dioxins, coplanar polychlorinated biphenyls (PCBs), parabens hydroxybenzoate derivatives, bisphenol A (BPA), polycyclic aromatic hydrocarbons (PAHs), organotins) used as industrial solvents/lubricants, plasticizers, pharmaceutical agents, and flame retardants, and natural substances (e.g., estrone (E1), 17β-estradiol (E2), and estriol (E3); natural androgens e.g., testosterone; phytosteroids e.g., β-sitosterol; isoflavonoids e.g., daidzeine) [18,27,28,29,30,31]. The chemical structures of the commonly recognized EDCs are shown in Figure 1.
The most frequently studied endocrine disruptors are pesticides, bisphenols, phthalates, synthetic and natural hormones, and polychlorinated biphenyls [32,33,34]. They are generally found in the order of nanograms to micrograms per liter (ng/L and μg/L) in the environment and can be identified in water using chemical analytical methods, like high-performance liquid chromatography and gas chromatography with mass spectroscopy, and biological methods or biosensors [35,36]. Commonly used pesticides acting as EDCs include organophosphorus pesticides (OPPs), organochlorine pesticides (OCPs) and herbicides such as chlorotriazine (CTs) and glyphosate [37]. OPPs are widely employed in agriculture (e.g., chlorpyrifos with 45,000 tons produced in 2016) and their residues in all food products and related environments need to be estimated in order to assess the risks of human exposure [38]. Atrazine is one of the most popular CTs herbicides used for many years and, together with other dangerous EDCs (equally banned in most countries) is again detected in surface and underground water all over the world [39]. Although the use of synthetic pesticides in agriculture has helped to increase food production, there is a great cost to human health, the environment, and its resources. Another widespread contaminant acting as EDC is bisphenol A, an artificial estrogen found in many hard plastics and hygiene products. It has been used to enhance the rapid growth of cattle and poultry and as an estrogen replacement for women [40].
In 2013, the World Health Organization highlighted that exposure to these chemicals is an issue of concern for wildlife and humans and that decision-makers need to take action to regulate human and environmental exposure to these chemicals [21].
EDCs can reach the aquatic environment through different pathways, e.g., wastewater discharge and release of pesticide residues from agricultural activities. Fish and wildlife can be directly exposed, and humans may become exposed through the intake of contaminated water and sea products. These substances, like CECs, are not easily removed from water through conventional treatment processes offered by water or sewage treatment plants. Thus, advanced removal technologies could represent more appropriate removal pathways. Owing to the diverse physicochemical properties of the endocrine disruptors, several processes can be applied as treatment technologies and obtain different removal efficiencies [41]. Data from ecological studies, animal models, clinical observations in humans, and epidemiological studies agree that endocrine-disrupting chemicals are significant for wildlife and human health [42,43].
Ultimately, EDCs are widespread in the environment and the increase of some effects such as diabetes, obesity, cognition deficits, neurodegenerative diseases, early puberty, thyroid dysfunction, heart diseases, and infertility have been hypothesized by some scientists to be linked to human exposure to these substances [44]. However, this relationship is highly debated in the literature and some authors have shown that links between EDCs and human health effects are weak to modest, so it needs to be further evaluated in future studies. Due to the increasing interest in this topic, this review reports an overview of the most important EDCs discovered in water, animals, and human exposure to them, analytical methods for their detection, and technologies useful to remove these compounds from waterbodies.

2. The Methodological Approach of the Review

In this review, the authors attempt to discuss the presence of EDCs in the aquatic compartment with particular attention to possible consequences due to human exposure to these compounds. Based on the scientific literature related to endocrine disruptors, this review has four main aims: (1) to summarize contamination and accumulation of emerging organic contaminants such as EDCs in water; (2) to outline and discuss the analytical methods used to assess EDC concentration, technical considerations and limitations; (3) to describe treatment technologies for the removal of these compounds from water environments; (4) to delineate the urgency and seriousness of EDC occurrence in water by emphasizing the different interactions with humans and health implications. The keywords “organic contaminants”, “environment”, “endocrine disruptor”, “emerging contaminants”, and “water” were selected individually or jointly to search for relevant information on the Web of Science, Scopus, and Google Scholar. Key literature published between 2004 and 2021 (up to January) were assimilated and analyzed.

3. EDCs in the Aquatic Environment

3.1. Contamination Sources and Paths into the Environment

Endocrine-disrupting compounds are involved in the water compartment contamination [45,46,47] of both surface water and groundwater [48], the marine environment [49,50,51,52], wastewater [53,54,55,56], and rivers and lakes [57,58,59,60]. In the last few years, EDCs reached the aquatic environment through various routes such as pharmaceutical and hospital waste disposal, wastewater treatment plants, leaching of chemicals used in industrial and household items (detergents and personal care products), and release of pesticide residues from agricultural activities (see Figure 2).
Water is a dynamic system and not a static location for accumulating contaminants, so in order to assess contamination issues some distribution mechanisms must be taken into account including sorption to sediments that can result as a long-term source. Flow in the natural water cycle is dominated by rain events. Precipitation leaches contaminants from buildings, streets, land surfaces, and agricultural fields (e.g., pesticides) or improperly disposed wastes and transports them into surface waters of local rivers, lakes, and reservoirs [61]. So rainwater runoff has also been identified as a source for contaminants [62]. In some cases, stormwater is conveyed in combined sewer systems that collect rainwater runoff, domestic sewage, and industrial wastewater [63]. During very rainy periods, the wastewater volume in these sewer systems can exceed their capacity, so excess wastewater (containing not only storm water but also untreated human and industrial waste, and toxic materials) is discharged into the surface water of nearby streams, rivers, or other water bodies as combined sewer overflows. Typically, water discarded by households and commercial users, collected in the sewer system, is treated by wastewater treatment plants in order to be discharged into surface waters. Furthermore, excess surface water and runoff (both due to the rain) are percolated into aquifers, bodies of permeable rock which can contain or transmit groundwater, and withdrawn during dry periods, increasing water reliability, protecting water quality, and providing treatment. In cities that rely on groundwater for their drinking water supply, aquifers are used in bank filtration to purify surface water [64]. Surface water and groundwater transport across the water cycle can provide natural water purification, often referred to as natural attenuation, e.g., processes such as dilution, sorption, volatilization, and chemical or biological degradation. In these ways, contaminant concentrations could be attenuated below the detection limits [61]. A considerable amount of pollutants are released from sewer leakages into the sewershed’s groundwater [65]. EDCs as pesticides reach the soil from rainwater or irrigation water washing and then they can infiltrate into ground and surface waters [61]. Moreover, contaminants not totally removed by wastewater treatment, and so present in treated wastewater, are released into the receiving surface waters, resulting in long-term chronic exposure of the aquatic ecosystem [66].
EDCs’ presence in the environments and their consequent exposure risk have been analyzed only in recent years thanks to the application of appropriate and sensitive methods for their detection. The application of advanced chromatography and mass spectrometry technologies to environmental analysis has allowed the determination of a broader range and a more comprehensive assessment of environmental contaminants [67].
In a recent review, Gonsioroski et al. [68] described the common endocrine-disrupting chemicals present in aquatic environments and their effects on the reproductive system. The review highlighted that chemical contamination in water has originated from byproducts formed during water disinfection processes, release from industry and livestock activity, or therapeutic drugs released into sewage including disinfection byproducts, fluorinated compounds, bisphenol A, phthalates, pesticides, and estrogens. Several studies reported the formation of EDCs during drinking water treatments and the association between exposure and increased risk of cancer development and adverse reproductive outcomes [69,70,71,72,73].

3.2. Occurrence in Water

In the following subsection, we describe the occurrence ad abundance of EDCs in different water matrices, starting from freshwater and estuaries to seawater, wastewater, and also drinking water. All the data analyzed are summarized in Table 1.
The occurrence of hormones in surface freshwater was reported in several African and European countries at different concentration levels. Estrone concentrations in the range of 0.1–69 ng/L were detected in France, the Czech Republic, Italy, Germany, Luxembourg, and Spain, whereas 0.23–13.7 ng/L of progesterone were reported in France and Hungary. Testosterone and estriol were instead found in concentrations of up to 3 ng/L and 2.38 ng/L respectively, in Italy and France and 17β-estradiol was discovered in the concentration range of 0.33–5 ng/L in Hungary and Luxembourg [74]. The highest hormone levels were detected in Africa, where the discharge of untreated domestic and animal farm wastewater is common [74]. The African concentrations are from 3000 to 20,000 times higher than in Europe, with ranges of 3310–15,700 ng/L for 17β-estradiol and 510–45,500 ng/L for estriol [75]. In Portugal, the presence of EDCs was investigated in different rivers (Minho, Ave, and Mondego) to evaluate their influence on the observed feminization phenomenon in male fish. Concentrations of estrogens were lower in surface water samples from Minho than Ave or Mondego estuaries, with the estrone as the main estrogen, followed by 17β-estradiol and 17α-ethynylestradiol. By converting estrogen concentrations in 17α-ethynylestradiol equivalents, the contribution of estrogens was 1.3 ng/L, 3.5 ng/L, and 2.4 ng/L, respectively, for Minho, Ave, and Mondego estuaries stressing out a high risk for local aquatic species. The concentrations of alkylphenols and alkylphenol ethoxylates at both Minho and Mondego estuaries were 600 ng/L and 2700 ng/L respectively, lower than those detected at the Ave (1070 ng/L and 4855 ng/L). This study also shows that in Ave and Mondego estuaries the concentrations of industrial estrogens were excessive, and the amount detected may induce an endocrine disruption in aquatic organisms [60]. Several pharmaceutical and antifungal residues (e.g., sulfamethoxazole) were discovered in 18 out of the 20 collected samples in the Romanian territory of the Danube River at concentrations ranging from 2.5 to 30 ng/L [59,88]. The occurrence of alkylphenols and bisphenol A was also observed in five estuaries along the Northwest coastal area of Spain, with maximal concentrations of 337 ng/L for NP and 146 ng L for BPA [76]. In Italy, the Higher Institute for Environmental Protection and Research (HIEPR) has identified several pesticides in surface and groundwater in the years 2017–2018. A higher percentage of detection was observed for surface water (25%) with respect to groundwater (15%) [89]. The substances that most often led to an overwhelming concern are the herbicide glyphosate and its metabolite aminomethylphosphonic acid (AMPA), metolachlor and its metabolite metolachlor-ESA, and the fungicides dimethomorph and azoxystrobin. The report from HIEPR highlighted the presence of a higher number of non-compliance cases in surface water (Figure 3) compared to the limits set by European legislation (2008/105/CE and 2013/39/UE) for these substances.
Moreover, the evolution of the EU legislation suggests a further reduction of the limits for many of these substances to increasingly guarantee a reduction in the risks associated with exposure. The new EQS (concentration thresholds below which no adverse impact on the medium occurs) to which reference will be made, in some cases, are extremely low and a considerable analytical effort will therefore be required in order to be sure to meet these limits [89].
Such hazardous presence of pollutants in estuaries indicates that terrestrial river input is an important source of EDCs to coastal and marine environments. Alkylphenols, alkylphenol ethoxylates and bisphenol A were detected in the seawater of Thermaikos Gulf, Northern Aegean Sea, Greece. Concentration ranges of 22–201 ng/L for NP and 10.6–52.3 ng/L for bisphenol A were observed, whereas steroid EDCs were not detected [52]. In Spain concentrations of NP up to 4100 ng/L have been measured near the Mediterranean coast [82], while lower concentrations were observed in the Catalonian coast, i.e., 210 ng/L of NP [83]. In a Portuguese coastal area, the concentrations of BPA and NP were in the range of 1.1–17 ng/L and 29–78 ng/L, respectively [77]. In the North Sea, NP was detected at a concentration range of 0.3–221 ng/L and BPA up to 249 ng/L [80], while lower concentrations of NP (1.3–21.3 ng/L) and BPA (0–5.7 ng/L) were discovered in the Baltic Sea of Germany [81]. In China, concentration levels of 0.98–43.7 ng/L for BPA and of 1.43 ng/L for E1 were detected in the East China Sea water [79], whereas natural hormones as E1 and E2, as well as the synthetic EE2, were detected in surface water samples on the northern shelf of the South China Sea near the Pearl River Estuary at concentrations of 1.1 ng/L, 0.7 ng/L, and 0.6 ng/L, respectively [84]. In the case of BPA, one of the chemicals that is the subject of regulatory decisions all over the world, concentrations in North American and European fresh and marine surface waters and sediments were monitored over the years of 1996–2014 [90]. The 95th percentile concentrations of BPA in freshwater were 0.30 µg/L both for North America and Europe, whereas those in marine water were 0.024 µg/L and 0.15 µg/L, respectively. Notwithstanding the increased production of BPA and polycarbonate plastic over the sampling period, BPA concentrations for both North America and Europe have not changed significantly. Moreover, concentrations of BPA in all samples did not exceed the EU Predicted No Effect Concentrations (1.5 μg/L and 0.15 μg/L for freshwater and marine organisms, respectively).
EDCs’ occurrence in wastewater was also investigated, with particular attention to the different contamination levels between effluent and influent of wastewater treatment plant (WWTP). For example, in Volos (Greece), pharmaceutical and personal care compounds belonging to classes (from antibiotics to disinfectants) were detected during a one year monitoring study on water samples collected from the influent and the effluent of a WWTP, reaching concentrations from 1 ng/L to 15,320 ng/L in the influents and between 18 ng/L and 9965 ng/L in the effluents [85], highlighting that most EDCs are not removed by the performed treatment. In addition, several pharmaceutical compounds were identified in samples from five WWTPs in Santorini (Aegean Sea, Greece) at concentrations of 0.6 ng/L for nordiazepam and 6822 ng/L for carbamazepine in the influent and a non-negligible amount (0.4 ng/L for 9-OH risperidone and 2200 ng/L for carbamazepine) in the effluent [56].
Moreover, some works reported in the literature investigated surface water, wastewater, and drinking water, discovering concentrations that may differ by also several orders of magnitude depending on water type. The concentration of 13 selected EDCs was monitored in untreated urban and industrial wastewater in Serbia with the aim to assess their impact on the Danube River basin and associated freshwaters employed as sources for drinking water [78]. Natural and synthetic estrogens were detected in surface and wastewater at concentrations ranging from 0.1 to 64.8 ng/L, but not in drinking water. In addition, total estrogenic activity surpassed the threshold of 1 ng/L of E2 in three surface water samples and over half of wastewater samples. Alkylphenols, instead, were present in concentration ranges of 1.1–78.3 ng/L in wastewater, 0.1–37.2 ng/L in surface water, and 0.4–7.9 ng/L in drinking water. Among all EDCs identified, bisphenol A was the most abundant in all water types, with detection frequencies of 84% in wastewater, 70% in surface water, and 57% in drinking water [78]. Recently, emerging pesticides (chlorpyrifos, carbendazim, atrazine, and some of their degradates) were detected in freshwater (lakes and rivers) and drinking water (tap and bottled water) in Vietnam. Concentration in lakes reached 86.7 ng/L for carbendazim and 49.3 ng/L for triazines, while in rivers triazines content increased to 164 ng/L. Furthermore, lower contamination was observed for drinking water with respect to freshwater with total pesticide concentrations of 39.3 ng/L and 3.54 ng/L found in tap water and bottled water respectively [39]. Very different concentration levels were instead observed for parabens in distinct water sources, in particular mg/L in the influents of WWTPs, μg/L in their effluents, and ng/L in drinking water [91].
In order to evaluate human exposure to endocrine disruptors, the difference between wastewater concentrations and environmental concentrations of contaminants must be taken into account. Considering the results discussed above, humans are potentially exposed to EDCs daily through drinking water intake. In fact, several EDCs have been identified in the treated drinking water supply worldwide [86,87,92], particularly in tap water in the concentration range of 0.2–5510 ng/L, with a maximum concentration (28,000 ng/L) detected in drinking water from the wells in India. Currently, the impact on human health derived from drinking water is considered negligible but the effects of long-term exposure (accumulation of EDCs also in mixtures up to the concentration of mg/L) are of concern [93]. Among EDCs all over the world, only a few of them have been regulated in the national drinking water standards, in particular BPA in Europe, the USA and Japan, NP in Japan, phthalates in China, Japan, and the USA, and E2 in Japan. Moreover, in some cases chlorination byproducts are more potent than their parent compounds, so chlorinated EDCs should also be taken into account in future drinking water regulations [94]. Greater scientific efforts should be devoted to monitoring EDCs in drinking water and the development of epidemiological studies in order to ensure safe access to drinking water.
In conclusion, endocrine disruptor compounds were detected in water environments all over the world in concentrations from ng/L to mg/L depending on the nature of the compound. Higher concentrations were discovered for chemicals from anthropogenic activities such as bisphenol A, nonylphenol, and phthalates, whereas natural hormones such E2 were detected at lower concentrations even if their activity is much higher than those of synthetic EDCs.

3.3. Analytical Methodologies for EDC Detection

The physical and chemical properties of EDCs influence instrument techniques, such as sensitivity and/or selectivity. An important parameter useful to compare the analytical methods is the limits of detection (LOD), calculated from the lowest analyte concentration producing a peak that could be reliably distinguished from the noise. So we reported the typical values of LODs for the most representative compounds for EDCs class i.e., bisphenol A (BPA), nonylphenol (NP), and 17β-estradiol (E2) achieved from the different analytical methods present in the literature. The main analytical techniques used for EDC detection are summarized in Table 2 with their strengths and limitations.
Gas chromatography coupled with a mass spectrometry detector (GC-MS) is one of the most powerful techniques for the separation and identification of organic pollutants, however, in the case of non-volatile compounds and polar molecules with one or more functional groups such as hydroxyl or carboxyl substituents (as is the case with the majority of EDCs), a derivatization step is often required before GC analysis. The derivatization step changes the properties of target molecules for better chromatographic separation and higher sensitivity of instrumental detection and prevents sample thermal decomposition. So, a disadvantage of the GC-MS analysis compared to liquid chromatography is the need for preparation sample processes that are often time consuming and may also introduce new interferences into the sample [97,98]. Recently, high-resolution gas chromatography-negative chemical ionization-mass spectrometry (HRGC-NCI-MS) was reported as an enhanced useful method to measure estrogens in the water sources thanks to the advantages related to the quickness, accuracy, and identification of complex chemical components [99]. Negative chemical ionization (NCI) is a soft ionization technique that provides high sensitivity and selectivity for compounds containing electronegative atoms or functional groups such as phenolic compounds [100]. Moreover, by introducing an electronegative group through derivatization, a large number of chemicals can be accurately and sensitively quantified using this approach. In the literature, the following rank order of methods for estrogen detection based on a number of studies has been reported: high-resolution gas chromatography-negative chemical ionization-mass spectrometry (HRGC-NCI-MS) (34.8%) > high-pressure liquid chromatography (HPLC) (30.5%) > liquid chromatography-high resolution mass spectrometry (LC-HRMS) (17.5%) > enzyme-linked immunosorbent assay (ELISA) (9.2%) > gas chromatography-mass spectrometry (GC-MS) (4%)~liquid chromatography-mass spectrometry (LC-MS) (4%) [101]. The HRGC-NCI-MS method is able to reach a LOD of nanograms per liter (ng/L) or slightly below (0.02 ng/L for BPA, 0.05 ng/L for NP, and 0.1 ng/L for E2) [102], which is very important in the case of phenols and steroids presents in the water at concentrations of ng/L and pg/L respectively.
Liquid chromatography methods (HPLC, LC-HRMS, and LC-MS), instead, eliminate the need for the derivatization step, which requires skillful analysts to obtain optimum results. However, the high capital cost involved for sophisticated instruments such as LC-tandem mass spectrometry (LC-MS-MS) makes the GC-MS an excellent alternative (typical LODs values: 1.5 ng/L for BPA, 0.3 ng/L for NP, and 0.1 ng/L for E2 [103]) for researchers with hands-on experience and skill in handling samples, providing comparable sensitivity and selectivity to the LC-MS [104]. Some works demonstrated that LC-MS-MS can examine and monitor a variety of organic pollutants consisting of a multi-class of EDCs compared to other chromatograms like HPLC and GC-MS. Due to the instrumental sensitivity and accuracy, LC-MS-MS permits identifying and quantifying the multi-class of targeted EDCs in environmental samples (LODs values: 5.7 ng/L for BPA, 2.7 ng/L for NP, and 3.3 ng/L for E2) that are expected to have a very low concentration (10–1000 ng/L) [105]. Furthermore, LC-MS combined with multiple reaction monitoring analysis is a sensitive tool for the qualitative and quantitative determination of the target analytes [106].
Although the above discussed chromatography techniques can identify and quantify EDCs with high sensitivity and accuracy, they also require expensive equipment, skilled technicians, and time-consuming sample preparation procedures [93,94,95]. For these reasons, alternative immuno-analytical methods for rapid detection of EDCs such as enzyme-linked immunosorbent assay (ELISA) and immunosensors were developed in the last years [107]. The ELISA method was successfully employed for the quantification of nonylphenol (LOD of 6 ng/L [95]), bisphenol A (LOD range 30–80 ng/L [96]), and hormones with good sensitivity (LOD of 0.2–5 ng/L for E2), accuracy (mean recovery 96%), and precision (RSD 7–10) [108]. These immuno-analytical methodologies have high sensitivity, cost effectiveness and simplification but employ bioantibodies that are unstable and require a long time to be prepared [109]. These disadvantages have been overcome by many researchers through synthesizing antibody alternatives able to imitate the molecular recognition characteristics of bioantibodies. For example, a sensitive plasmonic biomimetic-ELISA (PBELISA) method, which involves the use of molecularly imprinted polymers film as recognition elements and catalase-mediated growth of AuNPs as signal generation strategy, was recently reported in the literature for the detection of BPA [110]. This method has several remarkable advantages including time saving and cost effectiveness, and shows excellent selectivity and sensitivity for BPA with a LOD of 6.20 ng/L, making it suitable for the detection of trace BPA residues in real samples. However, the detection of BPA through ELISA is specially challenging because this method is less specific than others due to the fact that it can also detect other bisphenols [111,112].
In conclusion, both chromatographic techniques (GC-MS and LC-MS) and immuno-analytical methods (ELISA) were extensively reported in the literature for the detection EDCs in water matrices. In recent years the development of advanced analytical methodologies (e.g., HRGC-NCI-MS, LC-MS-MS, and PBELISA) have allowed reaching very low LOD for several EDCs (lower than ng/L) making their monitoring easier, but with a consequent increase of the cost of instrumentation and qualified personnel.

3.4. Removal of EDCs from Water

Current water and wastewater treatment technologies, including flocculation, chemical coagulation, precipitation, adsorption, membrane, and activated sludge processes, provide only the reduction of endocrine disruptor concentrations. In order to reach their total removal from waters, advanced processes need to be employed, thus minimizing the health issues that these compounds can cause even at low concentrations (below the LOAEL). Owing to the diverse physicochemical properties of the endocrine disruptors, several processes can be applied as treatment technologies, obtaining distinct removal efficiencies [41]. Different methods, like biosoption, adsorption, advanced oxidation, membrane filtration, and biodegradation, have been investigated as suitable treatment pathways [113,114,115,116] for the removal of EDCs, mainly bisphenol A, phthalates, natural and synthetic estrogens (estrone, 17α-ethynylestradiol, 17β-estradiol), parabens, alkylphenols, and pharmaceuticals.
Thanks to its exceptional characteristics such as great efficiency, low operative and maintenance costs, and no relevant byproduct generation, adsorption is one of the most employed processes for EDC removal from water. Several compounds such as clays [117], zeolites [118], biochars, bioadsorbents, metal-organic frameworks [119], graphene oxide [120], carbon nanotubes [121], and industrial waste have been studied as novel non-conventional adsorbents in alternative to the more expensive activated carbon (the most used material for micro-contaminant removal) in order to make the adsorption process for the treatment of endocrine disruptors more sustainable [122,123,124]. In this context, bioadsorbents can be considered as green and economical alternatives to commonly used carbon-based adsorbents thanks to the fact that they are available and abundant in nature. Consequently, biosorption methodology was widely used in drinking water and industrial wastewater treatment [125,126] because it is easy to apply, inexpensive, low energy consuming, and gives safe byproducts [55,127]. The most employed biosorbents are fungi, yeast, bacteria, algae, chitosan, wood, bio-polymers, and wastes of agriculture materials that are usually more selective than conventional adsorbents [128,129]. The possible use of a sulfonated derivative of coffee waste (CW-SO3H) as a convenient and effective biosorbent for the removal of BPA from an aqueous solution has been reported in the literature with the aim to obtain a biosorbent to clean contaminated water and reduce coffee waste at the same time [130]. CW-SO3H has shown good properties such as a calculated biosorption capacity of 270 mg/g for the removal of BPA, about five times higher than the commercially available activated carbon. Moreover, bean (Phaselous vulgaris) husk biomass residual wastes were recently employed to obtain activated carbon in the presence of orthophosphoric acid, reaching optimum results in the sequestration of ibuprofen (IBP) from aqueous solution at pH of 4.75 with a maximum monolayer adsorptive capacity of 50.00 mg/g at 50 °C [131].
Membrane filtration (microfiltration, ultrafiltration, nanofiltration, and reverse osmosis), instead, take advantage of the particular physicochemical characteristics of the material from which membrane is made to successfully reject a wide spectrum of endocrine disruptors [132,133,134]. Ultrafiltration removal efficiency can be improved by combining it with other technologies like advanced oxidation processes, chlorination, and ozonation. The ultrafiltration-ozonation hybrid system employed by Si et al. [135] was able to remove up to 99% of all endocrine disruptors under study (17β-estradiol, estriol, 17α-ethynylestradiol, and bisphenol A), compared to 46% and 70% reached using only ultrafiltration and ozonation, respectively. For the same reasons, membrane bioreactor is usually combined with nanofiltration or reverse osmosis process, thus improving the removal rate of contaminants such as bisphenol A, alkylphenols and carbamazepine [136,137].
Biological processes (both aerobic and anaerobic), such as activated sludge treatment [138], anaerobic digester systems [139], and fungal bioreactors [140] are also frequently used in the removal of endocrine disruptors, especially in combination with other tertiary treatments to improve their efficiency [132,141,142]. Through different enzymatic mechanisms (e.g., adsorption, accumulation and so on) bacteria, microalgae, and fungi can efficiently degrade EDCs with better performances for mixed populations with respect to individual microorganisms [143,144]. In this context, marine microalgae species such as P. globosa, N. oculata, D. salina, and P. subcordiformis can remove nonylphenol (NP) from polluted aquatic ecosystems via biosorption, biodegradation, or biotransformation, with efficiencies ranging from 43% to 91% [145]. Extracellular ligninolytic enzymes secreted by white rot fungi, instead, can biodegrade EDCs such as bisphenol A and nonylphenol with removal efficiency from wastewater in the range of 60–100% for BPA and 65–90% for NP depending on the fungi species, incubation time, and initial concentration of pollutants [146]. Among microorganisms, fungi gave the best results in terms of their ability to biologically remove EDCs from water [147] thanks to their very active enzymatic systems [144,148].
Regarding advanced oxidation processes (AOPs), several studies are reported in the literature for ozonation, UV/peroxide [91], Fenton, and photocatalysis [149], usually in combination to obtain better results. As an example, the UV-ozone combined process [150] is more active in the removal from the water of EDCs such as bisphenol A, 17β-estradiol, and estriol than ozone alone, while chlorine oxidant produces several byproducts without a reduction of the estrogenic activity of EDCs [151]. The Fenton process coupled with biological treatment using up-flow anaerobic sludge blanket reactor showed an initional EE2 removal (1000 μg/L of these compounds were spiked in the samples daily) of 99%, also reducing toxicity from 73% to 30% [152].
The application of EDC treatments is based on the various concentrations and complexity of EDC compounds, and it requires accurate and appropriate sampling, determination, extraction, quantification, storage, and preservation procedures. The individuation of an appropriate removal treatment could consider the EDC characteristics, in some cases each EDC contaminant has different treatment procedures. In practice, the membrane filtration process is an efficient method for EDC removal without further treatments. Nevertheless, like other EDC treatment methods, the membrane filtrations are not able to remove emerging contaminants completely. At the same time, adsorption may be an effective process at low cost and wide spectrum of reliability, even if the process requires sorbent regeneration or disposal there are sustainable solutions using waste products. On the other hand, advanced oxidation processes such as Fenton were successfully utilized to remove E3, BPA, diethylstilbestrol (DES), E2, and EE2 with removal efficiencies of 84.9%, 99.5%, 99.1%, 97.8%, and 84.5%, respectively as reported by Sun et al. [153]. The photo-Fenton process also showed a removal efficiency for pharmaceutical compounds, many types of hormones, phenolic, pesticide, and PPCP compounds, ranging from 95% to 100% [154,155,156]. Nevertheless, the limitation of these processes was determined by the infeasible regeneration of iron ions and the final treatment of effluent to meet the discharge standards for iron concentrations. Thus, the overall factor of limitations and challenges in EDC treatment methods such as solubility, hydrophilicity, degradability, and polarity influenced the applicability of the treatment techniques from the degradation pathways and the byproducts produced. In Table 3 we report a comparison of the most common methodologies useful for EDC removal from water, with attention to water sources and EDC types, highlighting the effectiveness and drawbacks of each treatment, providing a potential outlook of EDC treatment strategies in water and wastewater treatment systems.
A comparison of advanced oxidation processes for EDC removal was reported in the literature on the basis of multiple factors (such as process engineering, environmental, economic, and social parameters) and numerically scored (from 1 to 5 corresponding to descriptive variable of “very low” to “very high”) to describe the different performances. The results of this study highlighted that H2O2/O3 (perozonation) received the highest average ranking (3.45), and other processes showed comparable performance thanks to advantages related to an established technology, regulation, and public acceptance [157]. However, the Photo-Fenton process could be considered the best treatment for endocrine-disrupting compounds thanks to its technical characteristics and higher efficient removal of many different compounds such as hormones, phenolic compounds and pesticides.
In conclusion, since many EDCs are not degraded enough by the available microorganisms, biodegradation must be associated with other methods such as membrane filtrations and advanced oxidation processes to improve removal percentages. As a consequence, combined techniques are recommended for better utilization of the current treatment technologies in order to minimize the concentration of EDCs in water.

3.5. EDC Accumulation in Dynamic Systems

In this section, it has been shown that EDCs are persistent contaminants and are being detected throughout the water cycle (surface water, both freshwater and seawater, groundwater, rainwater runoff, wastewater, and drinking water) at different concentration levels. Accumulation of such pollutants in dynamic systems is difficult to assess, due to water transport across the water cycles. Despite the fact that natural attenuation can remove contaminants from the cycles, strong evidence of the persistence of EDCs in groundwater have been reported. For example, EDCs as sulfamethoxazole, 4-nonylphenol, 17β-estradiol, and pharmaceutical residues, have been detected in the same groundwater for decades, causing a long-term and probably sustained contamination of this water [158]. Moreover, sorption of these contaminants to sediments, which in some cases provides natural purification of surface water and groundwater, can result in a long-term source. Several examples are reported in the literature regarding EDC accumulation in the river and marine sediments [159,160,161]. Nonylphenol at a concentration of 22–645 µg/kg was detected in sediments of a Spanish river (the Llobregat River) [160] that receive waters from a sewage treatment plants, while in China 7.55–20.8 µg/kg of NP and 2.31–13.16 µg/kg of BPA were observed at the Pearl River Estuary [161] and 77–199 µg/kg of BPA in the Bahe River [162]. Furthermore, phtalates were detected at a concentration of 12–610 µg/kg in sediments of the Mediterranean Sea [159]. Due to the low biodegradability of the majority of EDCs, their persistence may affect soil biota and they can also be resuspended in water to be again in the water cycle.

4. Effects of Exposure to EDCs and Health Implications

4.1. EDCs’ Mode of Action

The effect of EDCs could be described in three different actions: endocrine activity [18]; deleterious and/or pathologic endocrine mediated activity [24]; the cause-effect relationship between substance and endocrine activity in exposed subjects [163,164,165]. The European Safety Authority (EFSA) proposes that EDCs disrupt the endocrine system by binding hormonal receptors and/or regulating genomic expression. Moreover, mechanisms involved in endocrine disruption seem to be related to epigenetic alterations, like histone modifications and methylation and/or acetylation of DNA [166,167]. Nuclear hormone receptors (NHRs) are the main receptors that are targeted by EDCs [168]. Thus, endocrine disruptors may mimic the natural hormone’s function, acting as agonists or antagonists of them [169,170,171]. NHRs are activated by steroid hormones and can induce long-term effects in their target cells. As an example, the estrogen receptor, androgen receptor, progesterone receptor, pregnane X receptor, constitutive androstane receptor, thyroid hormone receptor, retinoid X receptor, glucocorticoid receptor, and mineralocorticoid receptor belong to the NHRs family [172]. Ultimately, EDCs can bind to or interact with or activate hormone receptors; antagonize hormone receptors; alter hormone synthesis, receptor expression, transport across cell membranes, distribution levels and metabolism of hormones and signal transduction in hormone-responsive cells; induce epigenetic modifications in hormone-producing or hormone-responsive cells [173,174]. Figure 4 shows the schematic representation of the main EDC modes of action.
In recent literature, there is a heated debate concerning the presence or otherwise of phenomena such as the non-monotonic dose-response relationships (NMDRC) and low-dose effects of endocrine disruptors. The existence of these phenomena is very important because it has a significant impact on the way risk assessments are conducted for these chemicals. Concerning low-dose effects, low doses are considered those below the doses used for traditional toxicological studies or tested in traditional toxicology assessments or occurring in the range of human exposures [175]. The NMDRC is defined as a nonlinear relationship between dose and effect, occurring where the slope of the dose–response curve changes sign somewhere within the range of doses examined [175]. Traditional approaches in regulatory toxicology assume that the dose–response curve is always monotonic. Under this assumption, high-dose testing can be used as the standard for assessing chemical safety. However, if an NMDRC is present, there is no certainty that the lack of adverse effects at high doses also confirms safety at low doses and consequently possible effects at low doses cannot be predicted by those occurred at higher doses. For example, in the case of hormones, even moderate changes in concentration in the low-dose range can produce substantial changes in receptor occupancy and therefore generate significant changes in biological effects. Due to the non-monotonic behavior, concepts such as potency and threshold are difficult to assess for EDCs. In fact, under the hypothesis of a monotonic relationship, risks associated with hazards can be greatly reduced by decreasing exposure. Whereas, it might be necessary to eliminate the hazard entirely to ensure safety when an NMDR is observed because the reduction of exposure may have uncertain effects on risk [176]. Moreover, it makes it very difficult to predict a safe level of exposure and it cannot be assumed that there are thresholds below which EDC exposures are safe. This is the position of The Endocrine Society, supported by decades of endocrine science [177]. However, there is no consensus in the scientific community on the existence of these phenomena and their relevance in toxicology. In contrast to The Endocrine Society’s position, in fact, other authors assert that toxicology and biology predict that the threshold of adversity (defined as a position that separates dose levels at which the effect can occur from dose levels at which it will not occur) also exists at all endpoints for endocrine disruptors, even if it cannot be measured due to experimental science limits in determining the shape of the dose response at very low doses [23]. For these purposes, experimental values used in conventional toxicity testing for regulatory risk assessment are the NOAEL (no observed adverse effect level) defined as the highest dose at which there was not an observed adverse effect and the LOAEL (lowest observed adverse effect level) i.e., the lowest dose at which there was an observed adverse effect. The reference dose for toxicological studies strongly depends on the EDC type considered. Several epidemiology studies on EDCs indicate that harm is occurring in animals and humans that are exposed at or below the “safe” dose suggesting that the methods nowadays employed in regulatory toxicology must be improved to correctly predict EDCs contribution to human diseases or to identify the doses that can cause harm [178].
The dangerous effect of EDCs moves over a wide range of concentrations taking into account the heterogeneity in chemical structures of this class of compounds. The potency of a chemical to activate specific hormone receptors is determined by its affinity and efficacy. A compound with low affinity must have a good efficacy to bind the receptor site; thus a sufficient concentration is necessary to produce a cellular response [179]. Typically endogenous hormones have both higher affinity and higher efficacy than exogenous chemicals that make them very potent. For example, bisphenol A is reported to be 6 orders of magnitude less potent than 17β-estradiol [180]. As a consequence, considerably higher concentrations of synthetic EDCs with respect to natural hormones are required to attain sufficient receptor occupancy or to displace endogenous ligands, and so show observable adverse effects. For this reason, some authors assert that human exposure to synthetic EDCs is generally negligible as compared to natural ones that have higher endocrine activity [181]. However, in the case of phenolic compounds, such as alkylphenols and bisphenol A, concentrations (up to mg/L) detected in wastewater [182] were higher than the natural steroidal estrogen 17β-estradiol and the synthetic contraceptive 17α-ethinylestradiol that have higher estrogenic activity (ng/L) [183].

4.2. Transport of EDCs to Humans

Several papers considered the presence of EDCs effects in marine organisms: it is well known that fish, such as zebrafish, medaka and fathead minnow, food fish like carps (Cirrhinus mrigala, Catla, and Labeo rohita), followed by murrels (Channa marulius, C. punctatus and C. striatus), and catfish (Clarias gariepinus, C. batrachus and Heteropneustes fossilis) could be used as bioindicators to understand toxicity related to ingestion, sorption, bioaccumulation, and translocation, identifying the presence of EDCs [45,184,185].
EDCs bioaccumulate in the organisms via uptake of the chemicals directly from the environment (deconcentration) or by ingestion of other organisms containing the pollutants. Once they have entered a food chain, their concentrations tend to increase with trophic status (biomagnification), leading to the highest concentrations in the top predator. Moreover, biomagnification may also be due to increased body size and lipid contents, and higher metabolic activities.
Diet also plays an important role in the exposure to EDC through beverages. In China, in fact, several EDCs such as phthalates (14,400 ng/L), parabens, bisphenol analogs, benzophenone-type UV filters (20 ng/L), and triclosan (10 ng/mL) were discovered in 116 popular beverage samples. The results suggested that phthalates were the predominant EDCs in all beverages with daily intakes lower than their respective maximum acceptable doses suggested by various agencies, indicating a low potential health risk [186].
Based on the data reported above concerning drinking water contamination with EDCs, a possible route for human exposure is represented by drinking water. The World Health Organization (WHO) has proposed the following three representative EDCs for drinking water contamination and benchmark values: 0.01 μg/L for bisphenol A, 0.001 μg/L for 17β-estradiol, and 0.3 μg/L for nonylphenol. As indicated by the WHO, currently there is no evidence of risks to health from drinking water but these parameters have been included in the Directive on the basis of the precautionary principle since aquatic life is much more sensitive to the effects of estrogenic EDCs than humans [187]. In the last years in the US, the levels in drinking water have been reported to be at or below these numbers [188]. The US EPA (United States Environmental Protection Agency) did not propose restrictions for EDCs in the United States. Endocrine disruptors will be subjected by EPA to a comprehensive risk assessment by taking into account the differences between the levels of exposure that can produce adverse effects, and the typical exposure levels experienced by humans and wildlife in order to determine if this safety standard is appropriate to protect public health and the environment [189]. Although in 2018 the European Commission adopted a proposal for a revised drinking water directive that would add limit values for endocrine-disrupting chemicals to the list of criteria for monitoring water quality, it must be taken into account that humans and other organisms are often exposed to mixtures of chemical substances, the composition of which is not known and that the assessment scheme based on a single substance is not adequate.

4.3. Effect of EDCs: Animals, Humans, and Mixture Effect

The main concern related to endocrine disruptors is represented by the disorder they can activate in humans or wildlife by modifying the level of hormones in the body [190,191,192], given their mechanism of accumulation in tissues and the environment. In this subsection we report several studies on the effects observed in animals and humans due to exposure to EDCs.

4.3.1. Effect on Animals

In wildlife, EDCs are suspected in the decline of certain species (e.g., possible increased sterility in the American alligator), change of sex in fish and shellfish, and other problems [193,194,195,196]. In fact, some pesticides (thiocarbamates, chlororganics, imidazoles, triazoles, triazines) determine an antiandrogenic action, highlighted by the changes in macroscopic sexual findings in aquatic animals (particularly for exposure to herbicides and fungicides) such as the demasculinization in rats and fish [197], the production of estrogens and hermaphroditism in frogs [198], and other developmental disorders of the male gonad in alligators [199]. For example, a study on Daphnia Magna has shown that endosulfan sulfate disrupts the ecdysteroidal system (regulating processes such as molting and embryonic development) and juvenile hormone activity (regulating the sex ratio) of crustaceans [200,201]. Many studies have highlighted that EDCs can cause adverse effects in animals, including gene suppression and gene activation, even at very low concentrations (parts per billion and parts per trillion) [202,203,204]. Bosveld et al. reported some reproductive outcomes, hormone metabolism, and circulating steroid levels of fish-eating birds caused by effects of organohalogens [205]. Even in teleosts, seals, whales, and other mammals, several studies described the EDC-induced reproductive failure and thyroidal anomalies due to PCB and polybrominated diphenyl ether (PBDE) contamination [49,206,207]. Some experiments in zebra fish brain indicated that in the presence of diethylhexyl phthalate in the environment, particularly with very low doses, such as 0.02 mg/L, the organism could have a negative modulation of appetite stimulus, also confirmed by the real-time quantitative reverse transcription PCR analysis performed on key molecules involved in appetite control [208]. Moreover, pesticides such as dichlorodiphenyltrichloroethane (DDT) can block the receptors, inhibiting their function and the activity of hormones, thus altering hormonal feedback [209,210]. In many cases, the interaction of pesticides with the hormonal (endocrine) system of wildlife has led to impaired reproduction and gradual population decline of certain species [211]. It has several consequences on the endocrine system such as demasculinization and feminization of gonads [194] of male vertebrates, control of ovarian growth exerted by the neurohormones secreted at the eyestalk of the crab Neohelice granulata [212], and long-term histofunctional changes in the thyroid gland on the crocodilian specie of C. latirostris during embryonic development [213]. The endocrine-disrupting activity of these pesticides and their metabolites (e.g., dialkyl phosphates) has been observed in different animals. In particular, it was observed a homeostasis imbalance of hormones associated with the hypothalamus-pituitary-gonad axis [214] and hypothalamus-pituitaryadrenal axis [215] in rats, and with the hypothalamus-pituitary-thyroid axis in zebrafish [216].
The above reported effects are observed in species that are likely to be exposed to EDCs or that are used as a mammalian model for humans (e.g., rats). Most of them are related to sexual disorders that could have negative consequences in the preservation of animal species.

4.3.2. Effect on Humans

Several human organs and related glands are targeted by endocrine disruptors such as the brain (hypothalamus, pineal and pituitary glands), thyroid and parathyroid, adrenal gland, thymus, pancreas and gonads (testes and ovaries) as shown in Figure 5.
EDCs cause antiandrogenic effects even in humans. Moreover they mimic the estrogenic action, as confirmed by both experiments in vivo and in vitro [217,218,219,220,221,222]. This great attention to the reproductive system underlines how it can represent a sentinel organ for environmental stresses. In fact, epidemiological data and the most recent studies report human semen as an important source of early biomarkers in comparison with blood for assessing the environmental impact on human health, confirming its high sensitivity [223,224,225].
Pesticides were tested for antagonism to a human androgen receptor (hAR) by highly sensitive transactivation assays using Chinese hamster ovary cells. Results have shown that 66 of 200 pesticides under study exhibited antiandrogenic activity [29]. Supporting evidence of the dangerous effects of EDCs on the endocrine system and their complicated regulation mechanisms have been widely reported in the literature [226]. Organochlorine pesticides, used worldwide for several decades, are characterized by bioaccumulation in the environment, especially in the food chain, through which they reach the human body, and are largely detected in population screenings often correlated with various diseases [227]. In fact, cross-sectional association between use of OCPs and risk of hypothyroidism and hyperthyroidism among female spouses (n = 16.529) in Iowa and North Carolina (USA) was reported in the Agricultural Health Study in 1993–1997. In Taiwan [228] cow milk and beef consumption as well as menstruation characteristics were significantly associated with several OCPs (in particular hexachlorocyclohexanes) residues in breast milk. Furthermore, a correlation between β-and γ-hexachlorocyclohexanes and the infertility diseases of Taiwanese women was observed, so dietary habits might affect the exposure to these EDCs. The use of the OCPs chlordane, the fungicides benomyl and maneb/mancozeb, and the herbicide paraquat was significantly associated with thyroid dysfunctions such as hypothyroidism and hyperthyroidism (both solely for maneb/mancozeb). These results suggest a synergic role of organochlorines with fungicides in the development of thyroid diseases among women [229]. Different epidemiological studies have been conducted on the possible association between glyphosate (widely employed herbicides) exposure and the high risk of adverse reproductive outcomes and birth defects in the progeny, showing that women exposed to glyphosate increase the risk of late miscarriages and a decrease in fecundability. Moreover, this exposure during pregnancy is associated with an increase of anogenital distance in both males and females, increased testosterone levels in the female and several disturbances of developmental and reproductive parameters in progeny, such as retardation of the fetal skeleton [230]. Other EDCs such as phthalates were characterized by reproductive toxicity in humans and animals, many of them were antiandrogenic and they can cause infertility and reproductive problems in males [231]. They are more toxic to young children, who are much more susceptible to exposure, including fetal life [232]. A statistically significant and negative association between exposure to phthalates (such as monoethyl phthalate, monobutyl phthalate, and monobenzyl phthalate) and anogenital distance (AGD), penis length and width was observed in Mexican male newborns [233]. Concerning bisphenol A, it has several harmful effects on human health, especially for women. Thanks to its estrogenic activity, bisphenol A can bind to α-and β-estrogen receptors and act as a reproductive toxicant and affect fertility even at very low concentrations [234,235]. In fact, in a recent study, bisphenol A was detected in all infertile women with polycystic ovary syndrome [236]. Current research reveals that endocrine disruptors also interfere with energy metabolic homeostasis. Indeed, these compounds may decrease basal metabolic rate, modificate the regulation of appetite and satiety, impair adipose tissue by enhancing the number and size of adipose cells and varying their endocrine regulation and adipocytokine production [237]. An exposure to these chemicals (called obesogens) during early development increases such adverse health consequences, in addition to a predisposition to weight gain despite a correct balance of diet and physical activity [190,238]. Another target of endocrine-disrupting compounds is represented by the diencephalic system. In fact, besides their interaction with endocrine receptors, EDCs may mimic neurotransmitter actions, thus altering the proper function of the central nervous system [239]. Moreover, several EDCs, including bisphenol A, phthalates, polychlorinated biphenyls, and organochlorine pesticides, influence diabetes pathogenesis, and differential exposure to them might also contribute to racial/ethnic and economic inequality [240]. A correlation between exposure to endocrine disruptors and attention-deficit hyperactivity disorder, autism spectrum disorder, intellectual disability, global developmental delay, communication disorders, and neurodevelopmental disorders are reported by several studies. In particular, some EDCs, such as polybrominated diphenyl ethers, hexachlorobenzene, and bisphenol A, are reported to be a serious risk factor for the onset of neurodevelopmental disorders [241]. Furthermore, maternal exposure to plastic-derived endocrine-disrupting compounds (e.g., bisphenols, bis (2-ethylhexyl) phthalate and dibutyl phthalate) during the early development stages may affect pregnancy outcomes by altering embryo and placental development [242]. In fact, EDCs can produce more deleterious effects if exposure occurs during early development, referred as the time frame in which hormones are controlling cell changes to form tissues and organs. Exposure to EDCs during the developmental programming of a tissue, at lower doses than are required for effects in adults, could lead to changes in tissue development with effects that, although not apparent at birth, may appear later in life. However, these harmful consequences on the fetus development and postnatal health need to be further investigated by long-term studies. Typically the concentrations of EDCs found in the aquatic environment are lower than those at which these chemicals are considered harmful by regulatory toxicology. Nevertheless, for these particular chemicals, low doses could also have dangerous effects. Based on these considerations, research on effects of long-term exposures to low doses of EDCs should be performed in order to better evaluate the implications for human health.

4.3.3. Additive or Synergistic Effects of EDC Blends: Mixture Effects

In 2017 Barouki et al. [243] reported the additive or synergistic effects of EDCs; although a single compound could be innocuous, the combination of several endocrine disruptors might have dangerous consequences (i.e., the cocktail effect). The issue of blends has to be taken into consideration in the study of EDCs. The risk assessment, in fact, in the traditional scheme considers the effects of individual substances and does not take into account the possible effects of the mixtures present in the environment. In addition, due to the presence of mixtures, there is an awareness, at a scientific and regulatory level, that the risk deriving from these chemicals is underestimated. More attention and insights concerning the effects of chemical multi-exposure are hoped for by the authorities of the European Union. For this reason, particular caution is required even towards the lowest concentration levels. On the one hand, this is intended to be a critical reflection for the benefit of the experts because of the necessity of scientific insights, and on the other, for the legislators and administrators that can achieve increasingly sustainable management of the environment. However, the risk correlated to EDC mixtures is hard to assess due to the complexity of overall toxicity, their interaction, and the attendance of a sensitive period [181,244,245]. To improve the assessment of human exposure to chemical mixtures, researchers can apply in vitro assays to analyze the human health effects of these mixtures in biological samples. Moreover, the combination of in vitro assays with advanced high-resolution analytical tools can be advantageous to assess exposure and effect in humans due to mixtures of chemicals of concern and consequently identify chemicals that should be considered by regulatory authorities [246,247]. Definitely, the hypothesized causal relationship between EDC exposure and endocrine diseases needs to be further verified by long-term studies carried out on a wider number of subjects.

4.4. Strategies for the Reduction of EDC Pollution

Considering the above reported effect of EDC pollution, some strategies for the reduction of their sources and paths into the environment could be employed. Drinking water supplies must be protected from EDC contamination through tighter controls on sources and efficient treatment technologies. The discharge of chemicals into the sewer system must be banned and monitored in order to identify individual sources of contaminants. The same approach should be adopted for controlling the improper disposal of pharmaceutical and hospital waste and leaching of chemicals used in industrial and household items. Inclusion of limits for emerging contaminants such as EDCs in industrial wastewater discharges could lead to the reduction of the amount of harmful chemicals released to municipal wastewater by companies [248]. Moreover, several compounds with higher endocrine-disrupting potency should be replaced by environmentally friendly alternatives. This is especially relevant for pharmaceuticals and personal care products, plastic-derived compounds (phthalates and bisphenols), and pesticides. The use of some pesticides or antibiotics in animal husbandry has to be strongly restricted in all countries. Higher removal efficiencies must be reached for EDC removal in particular for heavily polluted wastewater and drinking water. Great attention must be paid to minimize the formation of treatment byproducts, notably in the case of chlorination processes due to the fact that chlorinated EDCs are often more potent than their parent compounds, e.g., in the case of BPA [94]. In conclusion, more appropriate limits for EDC concentration in water have to be proposed to ensure human safety, based on future studies on human health effects of long-term low-dose exposure to those EDCs that reach the human body through the consumption of water.

5. Conclusions

Endocrine disruptors have accumulated in the aquatic environment with consequent potential adverse exposure effects on humans such as homeostasis of the endocrine axis that leads to neurological, developmental, immunological, and reproductive disorders. Some chemicals that fall into this category are sex-steroid hormone mimics, pesticides, and fertilizers, derived from the waste of industries, agriculture, pharmaceutics, and sewage treatment plants. In this review, we analyzed the most recent literature related to EDCs in the environment, particularly in water, analytical methods, removal techniques and their potential human exposure routes and health implications. The key learning points of this review are contamination of water (freshwater, seawater, wastewater and drinking water); strengths and limitations of current analytical methods employed in endocrine disruptors detection; advantages and disadvantages of treatment technologies useful for EDC removal from water; biological effects on animals and humans. Some EDCs as hormones (E1, E2, E3, EE2), carbendazim, chlorpyrifos, dimethoate, iprodione, methomyl, tebuconazole, alkyl phenols, phthalates, and bisphenol A are almost ubiquitous in significant quanties in all the matrices, suggesting that they can easily enter our body. Most of the studies showed evidence of toxic effects in animals, which are part of the human diet. The analytical limits must, in particular, be adjusted to allow comparison with the EQS, which are often significantly lower, taking into account the provisions of the European Directives, which sets minimum efficiency criteria for the methods used to monitor the status of waters, sediments, and biota. EDC contamination is a complex phenomenon that is difficult to predict, both due to the large number of substances used and the multiplicity of paths they can follow in the environment. It must, therefore, be taken into account that humans and other organisms are often exposed to mixtures of chemical substances, the composition of which is not known a priori and that the assessment scheme based on a single substance is not adequate. It is necessary to take note of this evidence, confirmed worldwide, with a more cautious approach in the authorization phase. In order to better evaluate the implications for human health and confirm the hypothesized causal relationship between EDC exposure and endocrine diseases, studies on a wider number of subjects also exposed to low doses of EDCs for longer times should be performed.

Author Contributions

Conceptualization, L.M., O.M. and P.M.B.; methodology, C.P. and M.R.; validation, M.R., A.P. and L.M.; formal analysis, C.P. and M.R.; resources, O.M. and A.P.; writing—original draft preparation, M.R. and C.P.; writing—review and editing, C.P., M.R., P.M.B., L.M. and O.M.; visualization, C.P., M.R., A.P., P.M.B., L.M. and O.M.; supervision, L.M. and O.M. All authors have read and agreed to the published version of the manuscript.


This research received no external funding.

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.


This research has been funded by the University of Salerno (FARB 2017). The authors gratefully acknowledge Ylenia Miele for English language revision and Antonio Faggiano for graphical support.

Conflicts of Interest

The authors declare no conflict of interest.


  1. Patel, N.; Khan, M.D.; Shahane, S.; Rai, D.; Chauhan, D.; Kant, C.; Chaudhary, V.K. Emerging Pollutants in Aquatic Environment: Source, Effect, and Challenges in Biomonitoring and Bioremediation—A Review. Pollution 2020, 6, 99–113. [Google Scholar]
  2. Rosenfeld, P.E.; Feng, L.G. Risks of Hazardous Wastes; Elsevier: Amsterdam, The Netherlands, 2011. [Google Scholar]
  3. Richardson, S.D.; Ternes, T.A. Water Analysis: Emerging Contaminants and Current Issues. Anal. Chem. 2018, 90, 398–428. [Google Scholar] [CrossRef] [PubMed]
  4. Vigliotta, G.; Motta, O.; Guarino, F.; Iannece, P.; Proto, A. Assessment of perchlorate-reducing bacteria in a highly polluted river. Int. J. Hyg. Environ. Health 2010, 213, 437–443. [Google Scholar] [CrossRef] [PubMed]
  5. Iannece, P.; Motta, O.; Tedesco, R.; Carotenuto, M.; Proto, A. Determination of Perchlorate in Bottled Water from Italy. Water 2013, 5, 767–779. [Google Scholar] [CrossRef]
  6. Guarino, F.; Motta, O.; Turano, M.; Proto, A.; Vigliotta, G. Preferential Use of the Perchlorate over the Nitrate in the Respiratory Processes Mediated by the Bacterium Azospira sp. OGA24. Water 2020, 12, 2220. [Google Scholar] [CrossRef]
  7. Ricciardi, M.; Pironti, C.; Motta, O.; Miele, Y.; Proto, A.; Montano, L. Microplastics in the Aquatic Environment: Occurrence, Persistence, Analysis, and Human Exposure. Water 2021, 13, 973. [Google Scholar] [CrossRef]
  8. Fiorentino, A.; Rizzo, L.; Guilloteau, H.; Bellanger, X.; Merlin, C. Comparing TiO2 photocatalysis and UV-C radiation for inactivation and mutant formation of Salmonella typhimurium TA102. Environ. Sci. Pollut. Res. 2017, 24, 1871–1879. [Google Scholar] [CrossRef] [PubMed]
  9. Mortensen, A.; Granby, K.; Eriksen, F.D.; Cederberg, T.L.; Friis-Wandall, S.; Simonsen, Y.; Broesbøl-Jensen, B.; Bonnichsen, R. Levels and risk assessment of chemical contaminants in byproducts for animal feed in Denmark. J. Environ. Sci. Health Part B 2014, 49, 797–810. [Google Scholar] [CrossRef]
  10. Bilal, M.; Iqbal, H.M. An insight into toxicity and human-health-related adverse consequences of cosmeceuticals—A review. Sci. Total. Environ. 2019, 670, 555–568. [Google Scholar] [CrossRef]
  11. Kasonga, T.K.; Coetzee, M.A.; Kamika, I.; Ngole-Jeme, V.M.; Momba, M.N.B. Endocrine-disruptive chemicals as contaminants of emerging concern in wastewater and surface water: A review. J. Environ. Manag. 2021, 277, 111485. [Google Scholar] [CrossRef]
  12. Khan, N.A.; Khan, S.U.; Ahmed, S.; Farooqi, I.H.; Yousefi, M.; Mohammadi, A.A.; Changani, F. Recent trends in disposal and treatment technologies of emerging-pollutants—A critical review. TrAC Trends Anal. Chem. 2020, 122, 115744. [Google Scholar] [CrossRef]
  13. Motta, O.; Capunzo, M.; De Caro, F.; Brunetti, L.; Santoro, E.; Farina, A.; Proto, A. New approach for evaluating the public health risk of living near a polluted river. J. Prev. Med. Hyg. 2008, 49, 79–88. [Google Scholar]
  14. Proto, A.; Zarrella, I.; Capacchione, C.; Motta, O. One-Year Surveillance of the Chemical and Microbial Quality of Drinking Water Shuttled to the Eolian Islands. Water 2014, 6, 139–149. [Google Scholar] [CrossRef][Green Version]
  15. Pironti, C.; Motta, O.; Ricciardi, M.; Camin, F.; Cucciniello, R.; Proto, A. Characterization and authentication of commercial cleaning products formulated with biobased surfactants by stable carbon isotope ratio. Talanta 2020, 219, 121256. [Google Scholar] [CrossRef]
  16. Flint, S.; Markle, T.; Thompson, S.; Wallace, E. Bisphenol A exposure, effects, and policy: A wildlife perspective. J. Environ. Manag. 2012, 104, 19–34. [Google Scholar] [CrossRef]
  17. Jackson, J.; Sutton, R. Sources of endocrine-disrupting chemicals in urban wastewater, Oakland, CA. Sci. Total. Environ. 2008, 405, 153–160. [Google Scholar] [CrossRef] [PubMed]
  18. Zoeller, R.T.; Brown, T.R.; Doan, L.L.; Gore, A.C.; Skakkebaek, N.E.; Soto, A.M.; Woodruff, T.J.; Saal, F.S.V. Endocrine-Disrupting Chemicals and Public Health Protection: A Statement of Principles from The Endocrine Society. Endocrinology 2012, 153, 4097–4110. [Google Scholar] [CrossRef] [PubMed]
  19. Sharma, V.K.; Anquandah, G.A.K.; Nesnas, N. Kinetics of the oxidation of endocrine disruptor nonylphenol by ferrate(VI). Environ. Chem. Lett. 2009, 7, 115–119. [Google Scholar] [CrossRef]
  20. Colborn, T.; Saal, F.S.V.; Soto, A.M. Developmental effects of endocrine-disrupting chemicals in wildlife and humans. Environ. Health Perspect. 1993, 101, 378–384. [Google Scholar] [CrossRef] [PubMed]
  21. WHO. State of the Science of Endocrine Disrupting Chemicals. Available online: (accessed on 20 November 2020).
  22. Yilmaz, B.; Terekeci, H.; Sandal, S.; Kelestimur, F. Endocrine disrupting chemicals: Exposure, effects on human health, mechanism of action, models for testing and strategies for prevention. Rev. Endocr. Metab. Disord. 2019, 21, 127–147. [Google Scholar] [CrossRef]
  23. Brescia, S. Thresholds of adversity and their applicability to endocrine disrupting chemicals. Crit. Rev. Toxicol. 2020, 50, 213–218. [Google Scholar] [CrossRef] [PubMed]
  24. Gore, A.C.; Chappell, V.A.; Fenton, S.E.; Flaws, J.A.; Nadal, A.; Prins, G.S.; Toppari, J.; Zoeller, R.T. Executive Summary to EDC-2: The Endocrine Society’s Second Scientific Statement on Endocrine-Disrupting Chemicals. Endocr. Rev. 2015, 36, 593–602. [Google Scholar] [CrossRef][Green Version]
  25. Solecki, R.; Kortenkamp, A.; Bergman, Å.; Chahoud, I.; Degen, G.H.; Dietrich, D.; Greim, H.; Håkansson, H.; Hass, U.; Husoy, T.; et al. Scientific principles for the identification of endocrine-disrupting chemicals: A consensus statement. Arch. Toxicol. 2017, 91, 1001–1006. [Google Scholar] [CrossRef] [PubMed][Green Version]
  26. Wheeler, J.R.; Coady, K. Are all chemicals endocrine disruptors? Integr. Environ. Assess. Manag. 2016, 12, 402–403. [Google Scholar] [CrossRef][Green Version]
  27. Vinggaard, A.; Hnida, C.; Breinholt, V.; Larsen, J. Screening of selected pesticides for inhibition of CYP19 aromatase activity in vitro. Toxicol. Vitr. 2000, 14, 227–234. [Google Scholar] [CrossRef]
  28. Andersen, H.R.; Vinggaard, A.M.; Rasmussen, T.H.; Gjermandsen, I.M.; Bonefeld-Jørgensen, E.C. Effects of Currently Used Pesticides in Assays for Estrogenicity, Androgenicity, and Aromatase Activity in Vitro. Toxicol. Appl. Pharmacol. 2002, 179, 1–12. [Google Scholar] [CrossRef] [PubMed]
  29. Kojima, H.; Katsura, E.; Takeuchi, S.; Niiyama, K.; Kobayashi, K. Screening for estrogen and androgen receptor activities in 200 pesticides by in vitro reporter gene assays using Chinese hamster ovary cells. Environ. Health Perspect. 2004, 112, 524–531. [Google Scholar] [CrossRef][Green Version]
  30. Lemaire, G.; Mnif, W.; Pascussi, J.-M.; Pillon, A.; Rabenoelina, F.; Fenet, H.; Gomez, E.; Casellas, C.; Nicolas, J.-C.; Cavaillès, V.; et al. Identification of New Human Pregnane X Receptor Ligands among Pesticides Using a Stable Reporter Cell System. Toxicol. Sci. 2006, 91, 501–509. [Google Scholar] [CrossRef] [PubMed][Green Version]
  31. Lemaire, G.; Mnif, W.; Mauvais, P.; Balaguer, P.; Rahmani, R. Activation of α- and β-estrogen receptors by persistent pesticides in reporter cell lines. Life Sci. 2006, 79, 1160–1169. [Google Scholar] [CrossRef]
  32. Caliman, F.A.; Gavrilescu, M. Pharmaceuticals, Personal Care Products and Endocrine Disrupting Agents in the Environment—A Review. CLEAN Soil Air Water 2009, 37, 277–303. [Google Scholar] [CrossRef]
  33. Gore, A.C.; Crews, D.; Doan, L.L.; La Merrill, M.; Patisaul, H.; Zota, A. Introduction to Endocrine Disrupting Chemicals (EDCs)—A guide for public interest organizations and policy-makers. Endocr. Soc. J. 2014, 99, 21–22. [Google Scholar]
  34. Tijani, J.O.; Fatoba, O.O.; Babajide, O.O.; Petrik, L.F. Pharmaceuticals, endocrine disruptors, personal care products, nanomaterials and perfluorinated pollutants: A review. Environ. Chem. Lett. 2016, 14, 27–49. [Google Scholar] [CrossRef]
  35. Andrade-Eiroa, A.; Canle, M.; Leroy-Cancellieri, V.; Cerdà, V. Solid-phase extraction of organic compounds: A critical review (Part I). TrAC Trends Anal. Chem. 2016, 80, 641–654. [Google Scholar] [CrossRef]
  36. Gröger, T.M.; Käfer, U.; Zimmermann, R. Gas chromatography in combination with fast high-resolution time-of-flight mass spectrometry: Technical overview and perspectives for data visualization. TrAC Trends Anal. Chem. 2020, 122, 115677. [Google Scholar] [CrossRef]
  37. Rivoira, L.; Appendini, M.; Fiorilli, S.L.; Onida, B.; Del Bubba, M.; Bruzzoniti, M.C. Functionalized iron oxide/SBA-15 sorbent: Investigation of adsorption performance towards glyphosate herbicide. Environ. Sci. Pollut. Res. 2016, 23, 21682–21691. [Google Scholar] [CrossRef]
  38. Yang, F.-W.; Zhao, G.-P.; Ren, F.-Z.; Pang, G.-F.; Li, Y.-X. Assessment of the endocrine-disrupting effects of diethyl phosphate, a nonspecific metabolite of organophosphorus pesticides, by in vivo and in silico approaches. Environ. Int. 2020, 135, 105383. [Google Scholar] [CrossRef] [PubMed]
  39. Wan, Y.; Tran, T.M.; Nguyen, V.T.; Wang, A.; Wang, J.; Kannan, K. Neonicotinoids, fipronil, chlorpyrifos, carbendazim, chlorotriazines, chlorophenoxy herbicides, bentazon, and selected pesticide transformation products in surface water and drinking water from northern Vietnam. Sci. Total. Environ. 2021, 750, 141507. [Google Scholar] [CrossRef]
  40. Erler, C.; Novak, J. Bisphenol A Exposure: Human Risk and Health Policy. J. Pediatr. Nurs. 2010, 25, 400–407. [Google Scholar] [CrossRef] [PubMed]
  41. Gadupudi, C.K.; Rice, L.; Xiao, L.; Kantamaneni, K. Endocrine Disrupting Compounds Removal Methods from Wastewater in the United Kingdom: A Review. Science 2019, 1, 15. [Google Scholar] [CrossRef][Green Version]
  42. Myers, J.; Guillette, L.; Palanza, P.; Parmigiani, S.; Swan, S.; Saal, F.V. The Emerging Science of Endocrine Disruption. In International Seminar on Nuclear War and Planetary Emergencies? 30th Session; The Science and Culture Series? Nuclear Strategy and Peace Technology; World Scientific: Washington, DC, USA, 2004; pp. 105–121. [Google Scholar]
  43. Heindel, J.J.; Saal, F.S.V.; Blumberg, B.; Bovolin, P.; Calamandrei, G.; Ceresini, G.; Cohn, B.A.; Fabbri, E.; Gioiosa, L.; Kassotis, C.; et al. Parma consensus statement on metabolic disruptors. Environ. Health 2015, 14, 1–7. [Google Scholar] [CrossRef][Green Version]
  44. Trasande, L.; Zoeller, R.T.; Hass, U.; Kortenkamp, A.; Grandjean, P.; Myers, J.P.; DiGangi, J.; Bellanger, M.; Hauser, R.; Legler, J.; et al. Estimating Burden and Disease Costs of Exposure to Endocrine-Disrupting Chemicals in the European Union. J. Clin. Endocrinol. Metab. 2015, 100, 1245–1255. [Google Scholar] [CrossRef]
  45. Kar, S.; Sangem, P.; Anusha, N.; Senthilkumaran, B. Endocrine disruptors in teleosts: Evaluating environmental risks and biomarkers. Aquac. Fish. 2021, 6, 1–26. [Google Scholar] [CrossRef]
  46. Enachi, E.; Bahrim, G.E.; Ene, A.; de Jos, D. Pharmaceutical compounds and endocrine disruptors in aquatic environments: Ecotoxicological effects and analysis methodology. Ann. Dunarea de Jos Univ. Galati. Fascicle II Math. Physic Theor. Mech. 2019, 42, 172–182. [Google Scholar] [CrossRef]
  47. Rocha, M.J.; Rocha, E. Estrogenic Compounds in Estuarine and Coastal Water Environments of the Iberian Western Atlantic Coast and Selected Locations Worldwide—Relevancy, Trends and Challenges in View of the EU Water Framework Directive. Toxicol. Stud. Cells Drugs Environ. 2015. [Google Scholar] [CrossRef][Green Version]
  48. Gao, X.; Kang, S.; Xiong, R.; Chen, M. Environment-Friendly Removal Methods for Endocrine Disrupting Chemicals. Sustainability 2020, 12, 7615. [Google Scholar] [CrossRef]
  49. Krahn, M.M.; Hanson, M.B.; Schorr, G.S.; Emmons, C.K.; Burrows, D.G.; Bolton, J.L.; Baird, R.W.; Ylitalo, G.M. Effects of age, sex and reproductive status on persistent organic pollutant concentrations in “Southern Resident” killer whales. Mar. Pollut. Bull. 2009, 58, 1522–1529. [Google Scholar] [CrossRef] [PubMed][Green Version]
  50. Grieshaber, C.A.; Penland, T.N.; Kwak, T.J.; Cope, W.G.; Heise, R.J.; Mac Law, J.; Shea, D.; Aday, D.D.; Rice, J.A.; Kullman, S.W. Relation of contaminants to fish intersex in riverine sport fishes. Sci. Total Environ. 2018, 643, 73–89. [Google Scholar] [CrossRef]
  51. Müller, A.-K.; Markert, N.; Leser, K.; Kämpfer, D.; Crawford, S.E.; Schäffer, A.; Segner, H.; Hollert, H. Assessing endocrine disruption in freshwater fish species from a “hotspot” for estrogenic activity in sediment. Environ. Pollut. 2020, 257, 113636. [Google Scholar] [CrossRef]
  52. Arditsoglou, A.; Voutsa, D. Occurrence and partitioning of endocrine-disrupting compounds in the marine environment of Thermaikos Gulf, Northern Aegean Sea, Greece. Mar. Pollut. Bull. 2012, 64, 2443–2452. [Google Scholar] [CrossRef]
  53. Holmes, B.E.; Smeester, L.; Fry, R.C.; Weinberg, H.S. Identification of endocrine active disinfection by-products (DBPs) that bind to the androgen receptor. Chemosphere 2017, 187, 114–122. [Google Scholar] [CrossRef]
  54. Kar, S.; Senthilkumaran, B. Chapter 16—Water disinfection by-products cause acute toxicity in teleosts: A review. In Disinfection By-Products in Drinking Water; Prasad, M.N.V., Ed.; Butterworth-Heinemann: Oxford, UK, 2020; pp. 393–411. ISBN 978-0-08-102977-0. [Google Scholar]
  55. Bilal, M.; Rasheed, T.; Sosa-Hernández, J.E.; Raza, A.; Nabeel, F.; Iqbal, H.M.N. Biosorption: An Interplay between Marine Algae and Potentially Toxic Elements—A Review. Mar. Drugs 2018, 16, 65. [Google Scholar] [CrossRef][Green Version]
  56. Borova, V.L.; Maragou, N.C.; Gago-Ferrero, P.; Pistos, C.; Τhomaidis, Ν.S. Highly sensitive determination of 68 psychoactive pharmaceuticals, illicit drugs, and related human metabolites in wastewater by liquid chromatography–tandem mass spectrometry. Anal. Bioanal. Chem. 2014, 406, 4273–4285. [Google Scholar] [CrossRef][Green Version]
  57. Fox, G.A. What Have Biomarkers Told Us About the Effects of Contaminants on the Health of Fish-eating Birds in the Great Lakes? The Theory and a Literature Review. J. Great Lakes Res. 1993, 19, 722–736. [Google Scholar] [CrossRef]
  58. Chukwuka, A.; Ogbeide, O.; Uhunamure, G. Gonad pathology and intersex severity in pelagic (Tilapia zilli) and benthic (Neochanna diversus and Clarias gariepinus) species from a pesticide-impacted agrarian catchment, south-south Nigeria. Chemosphere 2019, 225, 535–547. [Google Scholar] [CrossRef] [PubMed]
  59. Bănăduc, D.; Rey, S.; Trichkova, T.; Lenhardt, M.; Curtean-Bănăduc, A. The Lower Danube River–Danube Delta–North West Black Sea: A pivotal area of major interest for the past, present and future of its fish fauna—A short review. Sci. Total. Environ. 2016, 545–546, 137–151. [Google Scholar] [CrossRef]
  60. Rocha, M.J.; Madureira, T.V.; Venade, C.S.; Martins, I.; Campos, J.; Rocha, E. Presence of estrogenic endocrine disruptors in three European estuaries in Northwest Iberian Peninsula (Portugal). Toxicol. Environ. Chem. 2019, 101, 244–264. [Google Scholar] [CrossRef]
  61. Pal, A.; He, Y.; Jekel, M.; Reinhard, M.; Gin, K.Y.-H. Emerging contaminants of public health significance as water quality indicator compounds in the urban water cycle. Environ. Int. 2014, 71, 46–62. [Google Scholar] [CrossRef] [PubMed]
  62. Bedoya-Ríos, D.F.; Lara-Borrero, J.; Duque-Pardo, V.; Madera-Parra, C.A.; Jimenez, E.M.; Toro, A.F. Study of the occurrence and ecosystem danger of selected endocrine disruptors in the urban water cycle of the city of Bogotá, Colombia. J. Environ. Sci. Health Part A 2017, 53, 317–325. [Google Scholar] [CrossRef]
  63. US EPA, R. 01 What Are Combined Sewer Overflows (CSOs)? Urban Environmental Program in New England. Available online: (accessed on 14 April 2021).
  64. Massmann, G.; Sültenfuß, J.; Dünnbier, U.; Knappe, A.; Taute, T.; Pekdeger, A. Investigation of groundwater residence times during bank filtration in Berlin: A multi-tracer approach. Hydrol. Process. 2008, 22, 788–801. [Google Scholar] [CrossRef]
  65. Bunting, S.; Lapworth, D.; Crane, E.; Grima-Olmedo, J.; Koroša, A.; Kuczyńska, A.; Mali, N.; Rosenqvist, L.; van Vliet, M.; Togola, A.; et al. Emerging organic compounds in European groundwater. Environ. Pollut. 2021, 269, 115945. [Google Scholar] [CrossRef] [PubMed]
  66. Musolff, A.; Leschik, S.; Reinstorf, F.; Strauch, G.; Schirmer, M. Micropollutant Loads in the Urban Water Cycle. Environ. Sci. Technol. 2010, 44, 4877–4883. [Google Scholar] [CrossRef]
  67. Fatta, D.; Achilleos, A.; Nikolaou, A.; Meriç, S. Analytical methods for tracing pharmaceutical residues in water and wastewater. TrAC Trends Anal. Chem. 2007, 26, 515–533. [Google Scholar] [CrossRef]
  68. Gonsioroski, A.; Mourikes, V.E.; Flaws, J.A. Endocrine Disruptors in Water and Their Effects on the Reproductive System. Int. J. Mol. Sci. 2020, 21, 1929. [Google Scholar] [CrossRef] [PubMed][Green Version]
  69. Hwang, B.-F.; Jaakkola, J.J.; Guo, H.-R. Water disinfection by-products and the risk of specific birth defects: A population-based cross-sectional study in Taiwan. Environ. Health 2008, 7, 23. [Google Scholar] [CrossRef] [PubMed][Green Version]
  70. Righi, E.; Bechtold, P.; Tortorici, D.; Lauriola, P.; Calzolari, E.; Astolfi, G.; Nieuwenhuijsen, M.J.; Fantuzzi, G.; Aggazzotti, G. Trihalomethanes, chlorite, chlorate in drinking water and risk of congenital anomalies: A population-based case-control study in Northern Italy. Environ. Res. 2012, 116, 66–73. [Google Scholar] [CrossRef]
  71. Villanueva, C.M.; Cantor, K.P.; Cordier, S.; Jaakkola, J.J.K.; King, W.D.; Lynch, C.F.; Porru, S.; Kogevinas, M. Disinfection Byproducts and Bladder Cancer: A Pooled Analysis. Epidemiology 2004, 15, 357–367. [Google Scholar] [CrossRef] [PubMed]
  72. Bove, G.E.; Rogerson, P.A.; Vena, J.E. Case control study of the geographic variability of exposure to disinfectant byproducts and risk for rectal cancer. Int. J. Health Geogr. 2007, 6, 18. [Google Scholar] [CrossRef][Green Version]
  73. Regli, S.; Chen, J.; Messner, M.J.; Elovitz, M.S.; Letkiewicz, F.J.; Pegram, R.A.; Pepping, T.; Richardson, S.D.; Wright, J.M. Estimating Potential Increased Bladder Cancer Risk Due to Increased Bromide Concentrations in Sources of Disinfected Drinking Waters. Environ. Sci. Technol. 2015, 49, 13094–13102. [Google Scholar] [CrossRef] [PubMed]
  74. Fekadu, S.; Alemayehu, E.; Dewil, R.; Van der Bruggen, B. Pharmaceuticals in freshwater aquatic environments: A comparison of the African and European challenge. Sci. Total. Environ. 2019, 654, 324–337. [Google Scholar] [CrossRef] [PubMed]
  75. Olatunji, O.S.; Fatoki, O.S.; Opeolu, B.O.; Ximba, B.J.; Chitongo, R. Determination of selected steroid hormones in some surface water around animal farms in Cape Town using HPLC-DAD. Environ. Monit. Assess. 2017, 189, 363. [Google Scholar] [CrossRef]
  76. Salgueiro-González, N.; Turnes-Carou, I.; Viñas-Diéguez, L.; Muniategui, S.; López-Mahía, P.; Prada-Rodríguez, D. Occurrence of endocrine disrupting compounds in five estuaries of the northwest coast of Spain: Ecological and human health impact. Chemosphere 2015, 131, 241–247. [Google Scholar] [CrossRef]
  77. Jonkers, N.; Sousa, A.; Galante-Oliveira, S.; Barroso, C.M.; Kohler, H.-P.E.; Giger, W. Occurrence and sources of selected phenolic endocrine disruptors in Ria de Aveiro, Portugal. Environ. Sci. Pollut. Res. 2009, 17, 834–843. [Google Scholar] [CrossRef] [PubMed][Green Version]
  78. Čelić, M.; Škrbić, B.D.; Insa, S.; Živančev, J.; Gros, M.; Petrović, M. Occurrence and assessment of environmental risks of endocrine disrupting compounds in drinking, surface and wastewaters in Serbia. Environ. Pollut. 2020, 262, 114344. [Google Scholar] [CrossRef] [PubMed]
  79. Shi, J.; Liu, X.; Chen, Q.; Zhang, H. Spatial and seasonal distributions of estrogens and bisphenol A in the Yangtze River Estuary and the adjacent East China Sea. Chemosphere 2014, 111, 336–343. [Google Scholar] [CrossRef] [PubMed]
  80. Heemken, O.; Reincke, H.; Stachel, B.; Theobald, N. The occurrence of xenoestrogens in the Elbe river and the North Sea. Chemosphere 2001, 45, 245–259. [Google Scholar] [CrossRef]
  81. Beck, I.-C.; Bruhn, R.; Gandrass, J. Analysis of estrogenic activity in coastal surface waters of the Baltic Sea using the yeast estrogen screen. Chemosphere 2006, 63, 1870–1878. [Google Scholar] [CrossRef]
  82. Petrovic, M.; Fernández-Alba, A.R.; Borrull, F.; Marce, R.M.; Mazo, E.G.; Barceló, D. Occurrence and distribution of nonionic surfactants, their degradation products, and linear alkylbenzene sulfonates in coastal waters and sediments in Spain. Environ. Toxicol. Chem. 2002, 21, 37–46. [Google Scholar] [CrossRef]
  83. González, S.; Petrovic, M.; Barceló, D. Simultaneous extraction and fate of linear alkylbenzene sulfonates, coconut diethanol amides, nonylphenol ethoxylates and their degradation products in wastewater treatment plants, receiving coastal waters and sediments in the Catalonian area (NE Spain). J. Chromatogr. A 2004, 1052, 111–120. [Google Scholar] [CrossRef]
  84. Deich, C.; Frazão, H.C.; Appelt, J.-S.; Li, W.; Pohlmann, T.; Waniek, J.J. Occurrence and distribution of estrogenic substances in the northern South China Sea. Sci. Total. Environ. 2021, 770, 145239. [Google Scholar] [CrossRef]
  85. Papageorgiou, M.; Kosma, C.; Lambropoulou, D. Seasonal occurrence, removal, mass loading and environmental risk assessment of 55 pharmaceuticals and personal care products in a municipal wastewater treatment plant in Central Greece. Sci. Total. Environ. 2016, 543, 547–569. [Google Scholar] [CrossRef]
  86. Lee, S.; Jeong, W.; Kannan, K.; Moon, H.-B. Occurrence and exposure assessment of organophosphate flame retardants (OPFRs) through the consumption of drinking water in Korea. Water Res. 2016, 103, 182–188. [Google Scholar] [CrossRef]
  87. Yang, G.C.; Yen, C.-H.; Wang, C.-L. Monitoring and removal of residual phthalate esters and pharmaceuticals in the drinking water of Kaohsiung City, Taiwan. J. Hazard. Mater. 2014, 277, 53–61. [Google Scholar] [CrossRef] [PubMed]
  88. Chitescu, C.L.; Kaklamanos, G.; Nicolau, A.I.; Stolker, A.A.M. (Linda) High sensitive multiresidue analysis of pharmaceuticals and antifungals in surface water using U-HPLC-Q-Exactive Orbitrap HRMS. Application to the Danube river basin on the Romanian territory. Sci. Total. Environ. 2015, 532, 501–511. [Google Scholar] [CrossRef]
  89. Rapporto Nazionale Pesticidi Nelle Acque Dati. 2017. Available online: (accessed on 22 January 2021).
  90. Staples, C.; van der Hoeven, N.; Clark, K.; Mihaich, E.; Woelz, J.; Hentges, S. Distributions of concentrations of bisphenol A in North American and European surface waters and sediments determined from 19 years of monitoring data. Chemosphere 2018, 201, 448–458. [Google Scholar] [CrossRef] [PubMed]
  91. Álvarez, M.A.; Ruidíaz-Martínez, M.; Cruz-Quesada, G.; López-Ramón, M.V.; Rivera-Utrilla, J.; Sánchez-Polo, M.; Mota, A.J. Removal of parabens from water by UV-driven advanced oxidation processes. Chem. Eng. J. 2020, 379, 122334. [Google Scholar] [CrossRef]
  92. Wee, S.Y.; Aris, A.Z. Endocrine disrupting compounds in drinking water supply system and human health risk implication. Environ. Int. 2017, 106, 207–233. [Google Scholar] [CrossRef] [PubMed]
  93. Wee, S.Y.; Aris, A.Z. Occurrence and public-perceived risk of endocrine disrupting compounds in drinking water. npj Clean Water 2019, 2, 1–14. [Google Scholar] [CrossRef][Green Version]
  94. Liu, Z.-H.; Dang, Z.; Liu, Y. Legislation against endocrine-disrupting compounds in drinking water: Essential but not enough to ensure water safety. Environ. Sci. Pollut. Res. 2021, 28, 1–6. [Google Scholar] [CrossRef]
  95. Céspedes, R.; Skryjová, K.; Raková, M.; Zeravik, J.; Fránek, M.; Lacorte, S.; Barceló, D. Validation of an enzyme-linked immunosorbent assay (ELISA) for the determination of 4-nonylphenol and octylphenol in surface water samples by LC-ESI-MS. Talanta 2006, 70, 745–751. [Google Scholar] [CrossRef] [PubMed]
  96. Lei, Y.; Fang, L.; Akash, M.S.H.; Liu, Z.; Shi, W.; Chen, S. Development and comparison of two competitive ELISAs for the detection of bisphenol A in human urine. Anal. Methods 2013, 5, 6106–6113. [Google Scholar] [CrossRef]
  97. Díaz-Cruz, M.S.; De Alda, M.J.L.; López, R.; Barceló, D. Determination of estrogens and progestogens by mass spectrometric techniques (GC/MS, LC/MS and LC/MS/MS). J. Mass Spectrom. 2003, 38, 917–923. [Google Scholar] [CrossRef]
  98. Mol, H.G.; Sunarto, S.; Steijger, O.M. Determination of endocrine disruptors in water after derivatization with N-methyl-N-(tert.-butyldimethyltrifluoroacetamide) using gas chromatography with mass spectrometric detection. J. Chromatogr. A 2000, 879, 97–112. [Google Scholar] [CrossRef]
  99. Spataro, F.; Ademollo, N.; Pescatore, T.; Rauseo, J.; Patrolecco, L. Antibiotic residues and endocrine disrupting compounds in municipal wastewater treatment plants in Rome, Italy. Microchem. J. 2019, 148, 634–642. [Google Scholar] [CrossRef]
  100. Yang, Y.; Lin, M.; Tang, J.; Ma, S.; Yu, Y. Derivatization gas chromatography negative chemical ionization mass spectrometry for the analysis of trace organic pollutants and their metabolites in human biological samples. Anal. Bioanal. Chem. 2020, 412, 6679–6690. [Google Scholar] [CrossRef]
  101. Matin, B.K.; Shakiba, E.; Moradi, M.; Zereshki, E.; Karami, A.; Vasseghian, Y.; Dragoi, E.-N.; Khaneghah, A.M. The concentration of estrogen in water resources: A systematic review and meta-analysis. Int. J. Environ. Anal. Chem. 2019, 1–10. [Google Scholar] [CrossRef]
  102. Kuch, H.M.; Ballschmiter, K. Determination of Endocrine-Disrupting Phenolic Compounds and Estrogens in Surface and Drinking Water by HRGC−(NCI)−MS in the Picogram per Liter Range. Environ. Sci. Technol. 2001, 35, 3201–3206. [Google Scholar] [CrossRef]
  103. Selvaraj, K.K.; Shanmugam, G.; Sampath, S.; Larsson, D.J.; Ramaswamy, B.R. GC–MS determination of bisphenol A and alkylphenol ethoxylates in river water from India and their ecotoxicological risk assessment. Ecotoxicol. Environ. Saf. 2014, 99, 13–20. [Google Scholar] [CrossRef]
  104. Omar, T.F.T.; Ahmad, A.; Aris, A.Z.; Yusoff, F.M. Endocrine disrupting compounds (EDCs) in environmental matrices: Review of analytical strategies for pharmaceuticals, estrogenic hormones, and alkylphenol compounds. TrAC Trends Anal. Chem. 2016, 85, 241–259. [Google Scholar] [CrossRef]
  105. Vega-Morales, T.; Sosa-Ferrera, Z.; Santana-Rodríguez, J. Determination of alkylphenol polyethoxylates, bisphenol-A, 17α-ethynylestradiol and 17β-estradiol and its metabolites in sewage samples by SPE and LC/MS/MS. J. Hazard. Mater. 2010, 183, 701–711. [Google Scholar] [CrossRef] [PubMed]
  106. Hao, C.; Zhao, X.; Yang, P. GC-MS and HPLC-MS analysis of bioactive pharmaceuticals and personal-care products in environmental matrices. TrAC Trends Anal. Chem. 2007, 26, 569–580. [Google Scholar] [CrossRef]
  107. Rochester, J.R. Bisphenol A and human health: A review of the literature. Reprod. Toxicol. 2013, 42, 132–155. [Google Scholar] [CrossRef]
  108. Manickum, T.; John, W. The current preference for the immuno-analytical ELISA method for quantitation of steroid hormones (endocrine disruptor compounds) in wastewater in South Africa. Anal. Bioanal. Chem. 2015, 407, 4949–4970. [Google Scholar] [CrossRef]
  109. Hong, S.; She, Y.; Cao, X.; Wang, M.; Zhang, C.; Zheng, L.; Wang, S.; Ma, X.; Shao, H.; Jin, M.; et al. Biomimetic enzyme-linked immunoassay based on a molecularly imprinted 96-well plate for the determination of triazophos residues in real samples. RSC Adv. 2018, 8, 20549–20556. [Google Scholar] [CrossRef][Green Version]
  110. Jia, M.; Chen, S.; Shi, T.; Li, C.; Wang, Y.; Zhang, H. Competitive plasmonic biomimetic enzyme-linked immunosorbent assay for sensitive detection of bisphenol A. Food Chem. 2020, 344, 128602. [Google Scholar] [CrossRef]
  111. Vandenberg, L.N.; Chahoud, I.; Heindel, J.J.; Padmanabhan, V.; Paumgartten, F.J.; Schoenfelder, G. Urinary, Circulating, and Tissue Biomonitoring Studies Indicate Widespread Exposure to Bisphenol A. Collect. Health Sci. 2012, 17, 407–434. [Google Scholar] [CrossRef]
  112. Ohkuma, H.; Abe, K.; Ito, M.; Kokado, A.; Kambegawa, A.; Maeda, M. Development of a highly sensitive enzyme-linked immunosorbent assay for bisphenol A in serum. Analyst 2001, 127, 93–97. [Google Scholar] [CrossRef] [PubMed]
  113. Okonkwo, J.O.; Sibali, L.L.; McCrindle, R.; Senwo, Z.N. An improved activated carbon method to quantify dichlorodiphenyltrichloroethane (DDT) in surface water. Environ. Chem. Lett. 2006, 5, 121–123. [Google Scholar] [CrossRef]
  114. Artham, T.; Doble, M. Bisphenol A and metabolites released by biodegradation of polycarbonate in seawater. Environ. Chem. Lett. 2012, 10, 29–34. [Google Scholar] [CrossRef]
  115. Li, G.; Zhang, X.; Sun, J.; Zhang, A.; Liao, C. Effective removal of bisphenols from aqueous solution with magnetic hierarchical rattle-like Co/Ni-based LDH. J. Hazard. Mater. 2020, 381, 120985. [Google Scholar] [CrossRef] [PubMed]
  116. Vieira, W.T.; De Farias, M.B.; Spaolonzi, M.P.; Da Silva, M.G.C.; Vieira, M.G.A. Removal of endocrine disruptors in waters by adsorption, membrane filtration and biodegradation. A review. Environ. Chem. Lett. 2020, 18, 1113–1143. [Google Scholar] [CrossRef]
  117. Djebri, N.; Boutahala, M.; Chelali, N.-E.; Boukhalfa, N.; Larbi, Z. Adsorption of bisphenol A and 2,4,5-trichlorophenol onto organo-acid-activated bentonite from aqueous solutions in single and binary systems. Desalination Water Treat. 2017, 66, 383–395. [Google Scholar] [CrossRef]
  118. Goyal, N.; Barman, S.; Bulasara, V.K. Quaternary ammonium salt assisted removal of genistein and bisphenol S from aqueous solution by nanozeolite NaY: Equilibrium, kinetic and thermodynamic studies. J. Mol. Liq. 2016, 224, 1154–1162. [Google Scholar] [CrossRef]
  119. Jun, B.-M.; Hwang, H.S.; Heo, J.; Han, J.; Jang, M.; Sohn, J.; Park, C.M.; Yoon, Y. Removal of selected endocrine-disrupting compounds using Al-based metal organic framework: Performance and mechanism of competitive adsorption. J. Ind. Eng. Chem. 2019, 79, 345–352. [Google Scholar] [CrossRef]
  120. Jiang, L.-H.; Liu, Y.-G.; Zeng, G.-M.; Xiao, F.-Y.; Hu, X.-J.; Hu, X.; Wang, H.; Li, T.-T.; Zhou, L.; Tan, X.-F. Removal of 17β-estradiol by few-layered graphene oxide nanosheets from aqueous solutions: External influence and adsorption mechanism. Chem. Eng. J. 2016, 284, 93–102. [Google Scholar] [CrossRef]
  121. Zhang, L.; Lv, J.; Xu, T.; Yang, L.; Jiang, X.; Li, Q. High efficiency removal and recovery of an endocrine disrupting compound–bisphenol AF from wastewaters. Sep. Purif. Technol. 2013, 116, 145–153. [Google Scholar] [CrossRef]
  122. De Andrade, J.R.; Oliveira, M.F.; da Silva, M.G.; Vieira, M.G. Adsorption of Pharmaceuticals from Water and Wastewater Using Nonconventional Low-Cost Materials: A Review. Ind. Eng. Chem. Res. 2018, 57, 3103–3127. [Google Scholar] [CrossRef]
  123. Maia, G.S.; de Andrade, J.R.; da Silva, M.G.; Vieira, M.G. Adsorption of diclofenac sodium onto commercial organoclay: Kinetic, equilibrium and thermodynamic study. Powder Technol. 2019, 345, 140–150. [Google Scholar] [CrossRef]
  124. De Souza, F.M.; Lazarin, A.M.; Vieira, M.G.A.; Dos Santos, O.A.A. Kinetic, equilibrium, and thermodynamic study on atrazine adsorption in organophilic clay. Desalin. Water Treat. 2018, 123, 240–252. [Google Scholar] [CrossRef]
  125. Coelho, C.M.; de Andrade, J.R.; da Silva, M.G.C.; Vieira, M.G.A. Removal of Propranolol Hydrochloride by Batch Biosorption Using Remaining Biomass of Alginate Extraction from Sargassum Filipendula Algae. Environ. Sci. Pollut. Res. 2020, 27, 16599–16611. [Google Scholar] [CrossRef]
  126. Sahu, O.; Singh, N. Significance of bioadsorption process on textile industry wastewater. In The Impact and Prospects of Green Chemistry for Textile Technology; Elsevier: Amsterdam, The Netherlands, 2019; pp. 367–416. [Google Scholar]
  127. Kyzas, G.Z.; Kostoglou, M.; Lazaridis, N.K.; Lambropoulou, D.A.; Bikiaris, D.N. Environmental friendly technology for the removal of pharmaceutical contaminants from wastewaters using modified chitosan adsorbents. Chem. Eng. J. 2013, 222, 248–258. [Google Scholar] [CrossRef]
  128. Crini, G.; Lichtfouse, E.; Wilson, L.D.; Morin-Crini, N. Conventional and non-conventional adsorbents for wastewater treatment. Environ. Chem. Lett. 2019, 17, 195–213. [Google Scholar] [CrossRef]
  129. Escudero, L.B.; Quintas, P.Y.; Wuilloud, R.G.; Dotto, G.L. Recent advances on elemental biosorption. Environ. Chem. Lett. 2019, 17, 409–427. [Google Scholar] [CrossRef]
  130. Ahsan, A.; Islam, T.; Imam, M.A.; Hyder, A.G.; Jabbari, V.; Dominguez, N.; Noveron, J.C. Biosorption of bisphenol A and sulfamethoxazole from water using sulfonated coffee waste: Isotherm, kinetic and thermodynamic studies. J. Environ. Chem. Eng. 2018, 6, 6602–6611. [Google Scholar] [CrossRef]
  131. Bello, O.S.; Alao, O.C.; Alagbada, T.C.; Olatunde, A.M. Biosorption of ibuprofen using functionalized bean husks. Sustain. Chem. Pharm. 2019, 13, 100151. [Google Scholar] [CrossRef][Green Version]
  132. Rodriguez-Narvaez, O.M.; Peralta-Hernandez, J.M.; Goonetilleke, A.; Bandala, E.R. Treatment technologies for emerging contaminants in water: A review. Chem. Eng. J. 2017, 323, 361–380. [Google Scholar] [CrossRef][Green Version]
  133. Bodzek, M.; Konieczny, K. Membranes in Organic Micropollutants Removal. Curr. Org. Chem. 2018, 22, 1070–1102. [Google Scholar] [CrossRef]
  134. Zielińska, M.; Cydzik-Kwiatkowska, A.; Bułkowska, K.; Bernat, K.; Wojnowska-Baryła, I. Treatment of Bisphenol A-Containing Effluents from Aerobic Granular Sludge Reactors with the Use of Microfiltration and Ultrafiltration Ceramic Membranes. Water Air Soil Pollut. 2017, 228, 282. [Google Scholar] [CrossRef] [PubMed][Green Version]
  135. Si, X.; Hu, Z.; Huang, S. Combined Process of Ozone Oxidation and Ultrafiltration as an Effective Treatment Technology for the Removal of Endocrine-Disrupting Chemicals. Appl. Sci. 2018, 8, 1240. [Google Scholar] [CrossRef][Green Version]
  136. Besha, A.T.; Gebreyohannes, A.Y.; Tufa, R.A.; Bekele, D.N.; Curcio, E.; Giorno, L. Removal of emerging micropollutants by activated sludge process and membrane bioreactors and the effects of micropollutants on membrane fouling: A review. J. Environ. Chem. Eng. 2017, 5, 2395–2414. [Google Scholar] [CrossRef]
  137. Kamaz, M.; Wickramasinghe, S.R.; Eswaranandam, S.; Zhang, W.; Jones, S.M.; Watts, M.J.; Qian, X. Investigation into Micropollutant Removal from Wastewaters by a Membrane Bioreactor. Int. J. Environ. Res. Public Health 2019, 16, 1363. [Google Scholar] [CrossRef][Green Version]
  138. Krah, D.; Ghattas, A.-K.; Wick, A.; Bröder, K.; Ternes, T.A. Micropollutant degradation via extracted native enzymes from activated sludge. Water Res. 2016, 95, 348–360. [Google Scholar] [CrossRef][Green Version]
  139. Yang, S.; Hai, F.I.; Price, W.E.; McDonald, J.; Khan, S.J.; Nghiem, L.D. Occurrence of trace organic contaminants in wastewater sludge and their removals by anaerobic digestion. Bioresour. Technol. 2016, 210, 153–159. [Google Scholar] [CrossRef] [PubMed][Green Version]
  140. Espinosa-Ortiz, E.J.; Rene, E.R.; Pakshirajan, K.; van Hullebusch, E.D.; Lens, P.N. Fungal pelleted reactors in wastewater treatment: Applications and perspectives. Chem. Eng. J. 2016, 283, 553–571. [Google Scholar] [CrossRef]
  141. Schmidt, N.; Page, D.; Tiehm, A. Biodegradation of pharmaceuticals and endocrine disruptors with oxygen, nitrate, manganese (IV), iron (III) and sulfate as electron acceptors. J. Contam. Hydrol. 2017, 203, 62–69. [Google Scholar] [CrossRef]
  142. Kasonga, T.K.; Coetzee, M.A.; Van Zijl, C.; Momba, M.N.B. Removal of pharmaceutical’ estrogenic activity of sequencing batch reactor effluents assessed in the T47D-KBluc reporter gene assay. J. Environ. Manag. 2019, 240, 209–218. [Google Scholar] [CrossRef]
  143. Xiong, J.-Q.; Kurade, M.B.; Jeon, B.-H. Can Microalgae Remove Pharmaceutical Contaminants from Water? Trends Biotechnol. 2018, 36, 30–44. [Google Scholar] [CrossRef]
  144. Roccuzzo, S.; Beckerman, A.P.; Trögl, J. New perspectives on the bioremediation of endocrine disrupting compounds from wastewater using algae-, bacteria- and fungi-based technologies. Int. J. Environ. Sci. Technol. 2021, 18, 89–106. [Google Scholar] [CrossRef]
  145. Wang, L.; Xiao, H.; He, N.; Sun, D.; Duan, S. Biosorption and Biodegradation of the Environmental Hormone Nonylphenol By Four Marine Microalgae. Sci. Rep. 2019, 9, 5277. [Google Scholar] [CrossRef]
  146. Grelska, A.; Noszczyńska, M. White Rot Fungi Can Be a Promising Tool for Removal of Bisphenol A, Bisphenol S, and Nonylphenol from Wastewater. Environ. Sci. Pollut. Res. 2020, 27, 39958–39976. [Google Scholar] [CrossRef] [PubMed]
  147. Liu, J.; Luo, Q.; Huang, Q. Removal of 17 β-estradiol from poultry litter via solid state cultivation of lignolytic fungi. J. Clean. Prod. 2016, 139, 1400–1407. [Google Scholar] [CrossRef]
  148. Da Silva, R.R. Bacterial and Fungal Proteolytic Enzymes: Production, Catalysis and Potential Applications. Appl. Biochem. Biotechnol. 2017, 183, 1–19. [Google Scholar] [CrossRef] [PubMed][Green Version]
  149. Varma, K.S.; Tayade, R.J.; Shah, K.J.; Joshi, P.A.; Shukla, A.D.; Gandhi, V.G. Photocatalytic degradation of pharmaceutical and pesticide compounds (PPCs) using doped TiO2 nanomaterials: A review. Water-Energy Nexus 2020, 3, 46–61. [Google Scholar] [CrossRef]
  150. Irmak, S.; Erbatur, O.; Akgerman, A. Degradation of 17β-estradiol and bisphenol A in aqueous medium by using ozone and ozone/UV techniques. J. Hazard. Mater. 2005, 126, 54–62. [Google Scholar] [CrossRef] [PubMed]
  151. Hu, J.-Y.; Aizawa, T.; Ookubo, S. Products of Aqueous Chlorination of Bisphenol A and Their Estrogenic Activity. Environ. Sci. Technol. 2002, 36, 1980–1987. [Google Scholar] [CrossRef]
  152. Frontistis, Z.; Xekoukoulotakis, N.P.; Hapeshi, E.; Venieri, D.; Fatta-Kassinos, D.; Mantzavinos, D. Fast degradation of estrogen hormones in environmental matrices by photo-Fenton oxidation under simulated solar radiation. Chem. Eng. J. 2011, 178, 175–182. [Google Scholar] [CrossRef]
  153. Sun, M.; Xu, D.; Ji, Y.; Liu, J.; Ling, W.; Li, S.; Chen, M. Using Fenton Oxidation to Simultaneously Remove Different Estrogens from Cow Manure. Int. J. Environ. Res. Public Health 2016, 13, 917. [Google Scholar] [CrossRef]
  154. Raji, M.; Mirbagheri, S.A.; Ye, F.; Dutta, J. Nano zero-valent iron on activated carbon cloth support as Fenton-like catalyst for efficient color and COD removal from melanoidin wastewater. Chemosphere 2021, 263, 127945. [Google Scholar] [CrossRef] [PubMed]
  155. Kohantorabi, M.; Giannakis, S.; Gholami, M.R.; Feng, L.; Pulgarin, C. A systematic investigation on the bactericidal transient species generated by photo-sensitization of natural organic matter (NOM) during solar and photo-Fenton disinfection of surface waters. Appl. Catal. B Environ. 2019, 244, 983–995. [Google Scholar] [CrossRef]
  156. Hu, C.; Huang, D.; Zeng, G.; Cheng, M.; Gong, X.; Wang, R.; Xue, W.; Hu, Z.; Liu, Y. The combination of Fenton process and Phanerochaete chrysosporium for the removal of bisphenol A in river sediments: Mechanism related to extracellular enzyme, organic acid and iron. Chem. Eng. J. 2018, 338, 432–439. [Google Scholar] [CrossRef]
  157. Fast, S.A.; Gude, V.G.; Truax, D.D.; Martin, J.; Magbanua, B.S. A Critical Evaluation of Advanced Oxidation Processes for Emerging Contaminants Removal. Environ. Process. 2017, 4, 283–302. [Google Scholar] [CrossRef]
  158. Barber, L.B.; Keefe, S.H.; LeBlanc, D.R.; Bradley, P.M.; Chapelle, F.H.; Meyer, M.T.; Loftin, K.A.; Kolpin, D.W.; Rubio, F. Fate of Sulfamethoxazole, 4-Nonylphenol, and 17β-Estradiol in Groundwater Contaminated by Wastewater Treatment Plant Effluent. Environ. Sci. Technol. 2009, 43, 4843–4850. [Google Scholar] [CrossRef][Green Version]
  159. Schmidt, N.; Castro-Jiménez, J.; Oursel, B.; Sempéré, R. Phthalates and organophosphate esters in surface water, sediments and zooplankton of the NW Mediterranean Sea: Exploring links with microplastic abundance and accumulation in the marine food web. Environ. Pollut. 2021, 272, 115970. [Google Scholar] [CrossRef] [PubMed]
  160. Petrovic, M.; Solé, M.; De Alda, M.J.L.; Barceló, D. Endocrine disruptors in sewage treatment plants, receiving river waters, and sediments: Integration of chemical analysis and biological effects on feral carp. Environ. Toxicol. Chem. 2002, 21, 2146–2156. [Google Scholar] [CrossRef]
  161. Diao, P.; Chen, Q.; Wang, R.; Sun, D.; Cai, Z.; Wu, H.; Duan, S. Phenolic endocrine-disrupting compounds in the Pearl River Estuary: Occurrence, bioaccumulation and risk assessment. Sci. Total. Environ. 2017, 584-585, 1100–1107. [Google Scholar] [CrossRef]
  162. Wang, X.; Chen, A.; Chen, B.; Wang, L. Adsorption of phenol and bisphenol A on river sediments: Effects of particle size, humic acid, pH and temperature. Ecotoxicol. Environ. Saf. 2020, 204, 111093. [Google Scholar] [CrossRef] [PubMed]
  163. Mnif, W.; Hassine, A.I.H.; Bouaziz, A.; Bartegi, A.; Thomas, O.; Roig, B. Effect of Endocrine Disruptor Pesticides: A Review. Int. J. Environ. Res. Public Health 2011, 8, 2265–2303. [Google Scholar] [CrossRef][Green Version]
  164. EFSA Scientific Committee. Scientific Opinion on the hazard assessment of endocrine disruptors: Scientific criteria for identification of endocrine disruptors and appropriateness of existing test methods for assessing effects mediated by these substances on human health and the environment. EFSA J. 2013, 11. [Google Scholar] [CrossRef]
  165. Slama, R.; Bourguignon, J.-P.; Demeneix, B.; Ivell, R.; Panzica, G.; Kortenkamp, A.; Zoeller, R.T. Scientific Issues Relevant to Setting Regulatory Criteria to Identify Endocrine-Disrupting Substances in the European Union. Environ. Health Perspect. 2016, 124, 1497–1503. [Google Scholar] [CrossRef][Green Version]
  166. Zama, A.M.; Uzumcu, M. Epigenetic effects of endocrine-disrupting chemicals on female reproduction: An ovarian perspective. Front. Neuroendocr. 2010, 31, 420–439. [Google Scholar] [CrossRef] [PubMed][Green Version]
  167. Baldi, F.; Mantovani, A. A new database for food safety: EDID (endocrine disrupting chemicals–diet interaction database). Reprod. Toxicol. 2008, 26, 57–63. [Google Scholar] [CrossRef]
  168. Combarnous, Y.; Nguyen, T.M.D. Comparative Overview of the Mechanisms of Action of Hormones and Endocrine Disruptor Compounds. Toxics 2019, 7, 5. [Google Scholar] [CrossRef][Green Version]
  169. Monneret, C. What is an endocrine disruptor? Comptes Rendus Biol. 2017, 340, 403–405. [Google Scholar] [CrossRef] [PubMed]
  170. Lubrano, C.; Genovesi, G.; Specchia, P.; Costantini, D.; Mariani, S.; Petrangeli, E.; Lenzi, A.; Gnessi, L. Obesity and Metabolic Comorbidities: Environmental Diseases? Oxidative Med. Cell. Longev. 2013, 2013, 1–9. [Google Scholar] [CrossRef][Green Version]
  171. Balaguer, P.; Delfosse, V.; Grimaldi, M.; Bourguet, W. Structural and functional evidences for the interactions between nuclear hormone receptors and endocrine disruptors at low doses. Comptes Rendus Biol. 2017, 340, 414–420. [Google Scholar] [CrossRef] [PubMed]
  172. Sever, R.; Glass, C.K. Signaling by Nuclear Receptors. Cold Spring Harb. Perspect. Biol. 2013, 5, a016709. [Google Scholar] [CrossRef][Green Version]
  173. Schug, T.T.; Janesick, A.; Blumberg, B.; Heindel, J.J. Endocrine disrupting chemicals and disease susceptibility. J. Steroid Biochem. Mol. Biol. 2011, 127, 204–215. [Google Scholar] [CrossRef][Green Version]
  174. De Coster, S.; Van Larebeke, N. Endocrine-Disrupting Chemicals: Associated Disorders and Mechanisms of Action. J. Environ. Public Health 2012, 2012, 1–52. [Google Scholar] [CrossRef]
  175. Vandenberg, L.N.; Colborn, T.; Hayes, T.B.; Heindel, J.J.; Jacobs, D.R., Jr.; Lee, D.-H.; Shioda, T.; Soto, A.M.; vom Saal, F.S.; Welshons, W.V. Hormones and Endocrine-Disrupting Chemicals: Low-Dose Effects and Nonmonotonic Dose Responses. Endocr. Rev. 2012, 33, 378–455. [Google Scholar] [CrossRef]
  176. Lagarde, F.; Beausoleil, C.; Belcher, S.M.; Belzunces, L.P.; Emond, C.; Guerbet, M.; Rousselle, C. Non-monotonic dose-response relationships and endocrine disruptors: A qualitative method of assessment. Environ. Health 2015, 14, 13. [Google Scholar] [CrossRef][Green Version]
  177. Demeneix, B.; Vandenberg, L.N.; Ivell, R.; Zoeller, R.T. Thresholds and Endocrine Disruptors: An Endocrine Society Policy Perspective. J. Endocr. Soc. 2020, 4, bvaa085. [Google Scholar] [CrossRef] [PubMed]
  178. Vandenberg, L.N. Low dose effects challenge the evaluation of endocrine disrupting chemicals. Trends Food Sci. Technol. 2019, 84, 58–61. [Google Scholar] [CrossRef]
  179. Borgert, C.J.; Baker, S.P.; Matthews, J.C. Potency matters: Thresholds govern endocrine activity. Regul. Toxicol. Pharmacol. 2013, 67, 83–88. [Google Scholar] [CrossRef] [PubMed][Green Version]
  180. Leffers, H.; Næsby, M.; Vendelbo, B.; Skakkebæk, N.E.; Jørgensen, M. Oestrogenic potencies of Zeranol, oestradiol, diethylstilboestrol, Bisphenol-A and genistein: Implications for exposure assessment of potential endocrine disrupters. Hum. Reprod. 2001, 16, 1037–1045. [Google Scholar] [CrossRef][Green Version]
  181. Autrup, H.; Barile, F.A.; Berry, S.C.; Blaauboer, B.J.; Boobis, A.; Bolt, H.; Borgert, C.J.; Dekant, W.; Dietrich, D.; Domingo, J.L.; et al. Human exposure to synthetic endocrine disrupting chemicals (S-EDCs) is generally negligible as compared to natural compounds with higher or comparable endocrine activity. How to evaluate the risk of the S-EDCs? Comput. Toxicol. 2020, 14, 100124. [Google Scholar] [CrossRef]
  182. Höhne, C.; Püttmann, W. Occurrence and temporal variations of the xenoestrogens bisphenol A, 4-tert-octylphenol, and tech. 4-nonylphenol in two German wastewater treatment plants. Environ. Sci. Pollut. Res. 2008, 15, 405–416. [Google Scholar] [CrossRef] [PubMed]
  183. Johnson, A.C.; Sumpter, J.P. Removal of Endocrine-Disrupting Chemicals in Activated Sludge Treatment Works. Environ. Sci. Technol. 2001, 35, 4697–4703. [Google Scholar] [CrossRef] [PubMed]
  184. Hutchinson, T.H.; Ankley, G.T.; Segner, H.; Tyler, C.R. Screening and Testing for Endocrine Disruption in Fish—Biomarkers As “Signposts,” Not “Traffic Lights,” in Risk Assessment. Environ. Health Perspect. 2006, 114, 106–114. [Google Scholar] [CrossRef] [PubMed][Green Version]
  185. Nagahama, Y.; Nakamura, M.; Kitano, T.; Tokumoto, T. Sexual plasticity in fish: A possible target of endocrine disruptor action. Environ. Sci. Int. J. Environ. Physiol. Toxicol. 2004, 11, 73–82. [Google Scholar]
  186. Wu, H.; Wu, L.-H.; Wang, F.; Gao, C.-J.; Chen, D.; Guo, Y. Several environmental endocrine disruptors in beverages from South China: Occurrence and human exposure. Environ. Sci. Pollut. Res. 2019, 26, 5873–5884. [Google Scholar] [CrossRef]
  187. Directive Review-Drinking Water-Environment-European Commission. Available online: (accessed on 22 February 2021).
  188. Seidel, C.J.; Samson, C.C.; Bartrand, T.; Ergul, A.; Summers, R.S.; Bartrand, T. Disinfection Byproduct Occurrence at Large Water Systems After Stage 2 DBPR. J. Am. Water Work. Assoc. 2017, 109, E287. [Google Scholar] [CrossRef][Green Version]
  189. Endocrine Disruptor Regulations and Lists in USA. Available online: (accessed on 19 February 2021).
  190. Heindel, J.J.; Newbold, R.R.; Schug, T.T. Endocrine disruptors and obesity. Nat. Rev. Endocrinol. 2015, 11, 653–661. [Google Scholar] [CrossRef]
  191. Nadal, A.; Quesada, I.; Tudurí, E.; Nogueiras, E.T.R.; Alonso-Magdalena, A.N.I.Q.P. Endocrine-disrupting chemicals and the regulation of energy balance. Nat. Rev. Endocrinol. 2017, 13, 536–546. [Google Scholar] [CrossRef] [PubMed]
  192. Braun, J.M. Early-life exposure to EDCs: Role in childhood obesity and neurodevelopment. Nat. Rev. Endocrinol. 2017, 13, 161–173. [Google Scholar] [CrossRef][Green Version]
  193. Hayes, T.B.; Collins, A.; Lee, M.; Mendoza, M.; Noriega, N.; Stuart, A.A.; Vonk, A. Hermaphroditic, demasculinized frogs after exposure to the herbicide atrazine at low ecologically relevant doses. Proc. Natl. Acad. Sci. USA 2002, 99, 5476–5480. [Google Scholar] [CrossRef] [PubMed][Green Version]
  194. Hayes, T.B.; Anderson, L.L.; Beasley, V.R.; de Solla, S.R.; Iguchi, T.; Ingraham, H.; Kestemont, P.; Kniewald, J.; Kniewald, Z.; Langlois, V.S.; et al. Demasculinization and feminization of male gonads by atrazine: Consistent effects across vertebrate classes. J. Steroid Biochem. Mol. Biol. 2011, 127, 64–73. [Google Scholar] [CrossRef][Green Version]
  195. Grilo, T.F.; Rosa, R. Intersexuality in aquatic invertebrates: Prevalence and causes. Sci. Total Environ. 2017, 592, 714–728. [Google Scholar] [CrossRef]
  196. Baatrup, E.; Junge, M. Antiandrogenic Pesticides Disrupt Sexual Characteristics in the Adult Male Guppy Poecilia Reticulata. Environ. Health Perspect. 2001, 109, 1063–1070. [Google Scholar] [CrossRef]
  197. Johnson, R.A.; Harris, R.E.; Wilke, R.A. Are pesticides really endocrine disruptors? WMJ Off. Publ. State Med. Soc. Wis. 2000, 99, 34–38. [Google Scholar]
  198. Guillette, L.J., Jr. Organochlorine Pesticides as Endocrine Disruptors in Wildlife. Cent. Eur. J. Public Health 2000, 8, 34–35. [Google Scholar]
  199. Guillette, L.J.; Gross, T.S.; Masson, G.R.; Matter, J.M.; Percival, H.F.; Woodward, A.R. Developmental abnormalities of the gonad and abnormal sex hormone concentrations in juvenile alligators from contaminated and control lakes in Florida. Environ. Health Perspect. 1994, 102, 680–688. [Google Scholar] [CrossRef]
  200. Palma, P.; Palma, V.; Matos, C.; Fernandes, R.; Bohn, A.; Soares, A.; Barbosa, I. Effects of atrazine and endosulfan sulphate on the ecdysteroid system of Daphnia magna. Chemosphere 2009, 74, 676–681. [Google Scholar] [CrossRef]
  201. Palma, P.; Palma, V.; Matos, C.; Fernandes, R.; Bohn, A.; Soares, A.; Barbosa, I. Assessment of the pesticides atrazine, endosulfan sulphate and chlorpyrifos for juvenoid-related endocrine activity using Daphnia magna. Chemosphere 2009, 76, 335–340. [Google Scholar] [CrossRef] [PubMed]
  202. Vom Saal, F.S.; Taylor, J.A.; Palanza, P.; Parmigiani, S. New Approaches to Risk Evaluation for Chemicals of Emerging Concern (CECs) That Have Endocrine Disrupting Effects. In Proceedings of the International Seminar on Nuclear War and Planetary Emergencies 44th Session, Erice, Italy, 19–24 August 2011; pp. 19–24. [Google Scholar]
  203. Saal, F.S.V. Triennial Reproduction Symposium: Environmental programming of reproduction during fetal life: Effects of intrauterine position and the endocrine disrupting chemical bisphenol A. J. Anim. Sci. 2016, 94, 2722–2736. [Google Scholar] [CrossRef]
  204. Parmigiani, S.; Saal, F.S.V.; Palanza, P.; Colborn, T.; Ragaini, R. Exposure to Very Low Doses of Endocrine Disrupting Chemicals (Edcs) During Fetal Life Permanently Alters Brain Development And Behavior In Animals And Humans. In Society and Structures; The Science and Culture Series? Nuclear Strategy and Peace Technology; World Scientific: Singapore, 2003; pp. 293–308. [Google Scholar] [CrossRef]
  205. Bosveld, A.T.; Berg, M.V.D. Reproductive failure and endocrine disruption by organohalogens in fish-eating birds. Toxicology 2002, 181–182, 155–159. [Google Scholar] [CrossRef]
  206. Nyman, M.; Koistinen, J.; Fant, M.L.; Vartiainen, T.; Helle, E. Current levels of DDT, PCB and trace elements in the Baltic ringed seals (Phoca hispida baltica) and grey seals (Halichoerus grypus). Environ. Pollut. 2002, 119, 399–412. [Google Scholar] [CrossRef]
  207. Ross, P.; Ellis, G.; Ikonomou, M.; Barrett-Lennard, L.; Addison, R. High PCB Concentrations in Free-Ranging Pacific Killer Whales, Orcinus orca: Effects of Age, Sex and Dietary Preference. Mar. Pollut. Bull. 2000, 40, 504–515. [Google Scholar] [CrossRef]
  208. Migliarini, B.; Piccinetti, C.; Martella, A.; Maradonna, F.; Gioacchini, G.; Carnevali, O. Perspectives on endocrine disruptor effects on metabolic sensors. Gen. Comp. Endocrinol. 2011, 170, 416–423. [Google Scholar] [CrossRef]
  209. Lemaire, G.; Terouanne, B.; Mauvais, P.; Michel, S.; Rahmani, R. Effect of organochlorine pesticides on human androgen receptor activation in vitro. Toxicol. Appl. Pharmacol. 2004, 196, 235–246. [Google Scholar] [CrossRef]
  210. Lee, H.-R.; Jeung, E.-B.; Cho, M.-H.; Kim, T.-H.; Leung, P.C.K.; Choi, K.-C. Molecular mechanism(s) of endocrine-disrupting chemicals and their potent oestrogenicity in diverse cells and tissues that express oestrogen receptors. J. Cell. Mol. Med. 2012, 17, 1–11. [Google Scholar] [CrossRef] [PubMed]
  211. Combarnous, Y. Endocrine Disruptor Compounds (EDCs) and agriculture: The case of pesticides. Comptes Rendus Biol. 2017, 340, 406–409. [Google Scholar] [CrossRef]
  212. Silveyra, G.R.; Canosa, I.S.; Zanitti, M.; Rodríguez, E.M.; Medesani, D.A. Interference of an atrazine commercial formulation with the endocrine control of ovarian growth exerted by the eyestalks. Environ. Sci. Pollut. Res. 2019, 27, 965–973. [Google Scholar] [CrossRef]
  213. Galoppo, G.H.; Tavalieri, Y.E.; Schierano-Marotti, G.; Osti, M.R.; Luque, E.H.; Muñoz-De-Toro, M.M. Long-term effects of in ovo exposure to an environmentally relevant dose of atrazine on the thyroid gland of Caiman latirostris. Environ. Res. 2020, 186, 109410. [Google Scholar] [CrossRef]
  214. Peiris, D.C.; Dhanushka, T. Low doses of chlorpyrifos interfere with spermatogenesis of rats through reduction of sex hormones. Environ. Sci. Pollut. Res. 2017, 24, 20859–20867. [Google Scholar] [CrossRef]
  215. Yang, F.-W.; Fang, B.; Pang, G.-F.; Ren, F.-Z. Organophosphorus pesticide triazophos: A new endocrine disruptor chemical of hypothalamus-pituitary-adrenal axis. Pestic. Biochem. Physiol. 2019, 159, 91–97. [Google Scholar] [CrossRef] [PubMed]
  216. Wu, S.; Li, X.; Liu, X.; Yang, G.; An, X.; Wang, Q.; Wang, Y. Joint toxic effects of triazophos and imidacloprid on zebrafish (Danio rerio). Environ. Pollut. 2018, 235, 470–481. [Google Scholar] [CrossRef]
  217. Singleton, D.W.; Khan, S.A. Xenoestrogen Exposure and Mechanisms of Endocrine Disruption. Front Biosci. 2003, 8, s110–s118. [Google Scholar] [CrossRef] [PubMed][Green Version]
  218. Akingbemi, B.T.; Hardy, M.P. Oestrogenic and antiandrogenic chemicals in the environment: Effects on male reproductive health. Ann. Med. 2001, 33, 391–403. [Google Scholar] [CrossRef]
  219. Sultan, C.; Balaguer, P.; Terouanne, B.; Georget, V.; Paris, F.; Jeandel, C.; Lumbroso, S.; Nicolas, J.-C. Environmental xenoestrogens, antiandrogens and disorders of male sexual differentiation. Mol. Cell. Endocrinol. 2001, 178, 99–105. [Google Scholar] [CrossRef]
  220. Hodges, L.C.; Bergerson, J.S.; Hunter, D.S.; Walker, C.L. Estrogenic effects of organochlorine pesticides on uterine leiomyoma cells in vitro. Toxicol. Sci. 2000, 54, 355–364. [Google Scholar] [CrossRef][Green Version]
  221. Quesada, I.; Fuentes, E.; Viso-León, M.C.; Soria, B.; Ripoll, C.; Nadal, A. Low Doses of the Endocrine Disruptor Bisphenol-A and the Native Hormone 17β-Estradiol Rapidly Activate the Transcription Factor CREB. FASEB J. 2002, 16, 1671–1673. [Google Scholar] [CrossRef]
  222. Takeuchi, T.; Tsutsumi, O. Serum Bisphenol A Concentrations Showed Gender Differences, Possibly Linked to Androgen Levels. Biochem. Biophys. Res. Commun. 2002, 291, 76–78. [Google Scholar] [CrossRef]
  223. Montano, L. Reproductive Biomarkers as Early Indicators for Assessing Environmental Health Risk. In Toxic Waste Management and Health Risk; Marfe, G., Di Stefano, C., Eds.; Bentham Science: Sharjah, United Arab Emirates, 2020; pp. 113–145. ISBN 978-981-14-5474-5. [Google Scholar]
  224. Montano, L.; Bergamo, P.; Volpe, M.G.; Lorenzetti, S.; Mantovani, A.; Notari, T.; Di Stasio, M.; Cerullo, S.; Cerino, P.; Iannuzzi, L. Human semen as an early, sensitive biomarker of environmental exposure: Preliminary results of the ECOFOODFERTILITY Project. Reprod. Toxicol. 2016, 64, 43–44. [Google Scholar] [CrossRef]
  225. Montano, L.; Bergamo, P.; Andreassi, M.G.; Vecoli, C.; Volpe, M.G.; Lorenzetti, S.; Mantovani, A.; Notari, T. The role of human semen for assessing environmental impact on human health in risk areas: Novels and early biomarkers of environmental pollution. EcoFoodFertility project. Reprod. Toxicol. 2017, 72, 44–45. [Google Scholar] [CrossRef]
  226. Lauretta, R.; Sansone, A.; Sansone, M.; Romanelli, F.; Appetecchia, M. Endocrine Disrupting Chemicals: Effects on Endocrine Glands. Front. Endocrinol. 2019, 10, 178. [Google Scholar] [CrossRef] [PubMed][Green Version]
  227. Martyniuk, C.J.; Mehinto, A.C.; Denslow, N.D. Organochlorine pesticides: Agrochemicals with potent endocrine-disrupting properties in fish. Mol. Cell. Endocrinol. 2020, 507, 110764. [Google Scholar] [CrossRef]
  228. Chen, M.-W.; Santos, H.M.; Que, D.E.; Gou, Y.-Y.; Tayo, L.L.; Hsu, Y.-C.; Chen, Y.-B.; Chen, F.-A.; Chao, H.-R.; Huang, K.-L. Association between Organochlorine Pesticide Levels in Breast Milk and Their Effects on Female Reproduction in a Taiwanese Population. Int. J. Environ. Res. Public Health 2018, 15, 931. [Google Scholar] [CrossRef][Green Version]
  229. Goldner, W.S.; Sandler, D.P.; Yu, F.; Hoppin, J.A.; Kamel, F.; LeVan, T.D. Pesticide Use and Thyroid Disease Among Women in the Agricultural Health Study. Am. J. Epidemiol. 2010, 171, 455–464. [Google Scholar] [CrossRef]
  230. Muñoz, J.P.; Bleak, T.C.; Calaf, G.M. Glyphosate and the key characteristics of an endocrine disruptor: A review. Chemosphere 2021, 270, 128619. [Google Scholar] [CrossRef] [PubMed]
  231. Balcı, A.; Özkemahlı, G.; Erkekoglu, P.; Zeybek, D.; Yersal, N.; Kocer-Gumusel, B. Effects of prenatal and lactational bisphenol a and/or di(2-ethylhexyl) phthalate exposure on male reproductive system. Int. J. Environ. Health Res. 2020, 1–14. [Google Scholar] [CrossRef] [PubMed]
  232. Przybylińska, P.A.; Wyszkowski, M. Environmental contamination with phthalates and its impact on living organisms. Ecol. Chem. Eng. S 2016, 23, 347–356. [Google Scholar] [CrossRef][Green Version]
  233. Bustamante-Montes, L.P.; Hernández-Valero, M.A.; García-Fàbila, M.; Halley-Castillo, E.; Karam-Calderón, M.A.; Borja-Aburto, V.H. Prenatal Phthalate Exposure and Decrease in Ano-Genital Distance in Mexican Male Newborns. Epidemiology 2008, 19, S270. [Google Scholar] [CrossRef]
  234. Nah, W.H.; Park, M.J.; Gye, M.C. Effects of early prepubertal exposure to bisphenol A on the onset of puberty, ovarian weights, and estrous cycle in female mice. Clin. Exp. Reprod. Med. 2011, 38, 75–81. [Google Scholar] [CrossRef][Green Version]
  235. Calaf, G.M.; Ponce‑Cusi, R.; Aguayo, F.; Bleak, T.C. Endocrine disruptors from the environment affecting breast cancer (Review). Oncol. Lett. 2020, 20, 19–32. [Google Scholar] [CrossRef][Green Version]
  236. Zhou, W.; Fang, F.; Zhu, W.; Chen, Z.-J.; Du, Y.; Zhang, J. Bisphenol A and Ovarian Reserve among Infertile Women with Polycystic Ovarian Syndrome. Int. J. Environ. Res. Public Health 2016, 14, 18. [Google Scholar] [CrossRef] [PubMed][Green Version]
  237. Street, M.E.; Angelini, S.; Bernasconi, S.; Burgio, E.; Cassio, A.; Catellani, C.; Cirillo, F.; Deodati, A.; Fabbrizi, E.; Fanos, V.; et al. Current Knowledge on Endocrine Disrupting Chemicals (EDCs) from Animal Biology to Humans, from Pregnancy to Adulthood: Highlights from a National Italian Meeting. Int. J. Mol. Sci. 2018, 19, 1647. [Google Scholar] [CrossRef][Green Version]
  238. Janesick, A.S.; Blumberg, B. Obesogens: An emerging threat to public health. Am. J. Obstet. Gynecol. 2016, 214, 559–565. [Google Scholar] [CrossRef] [PubMed][Green Version]
  239. Gore, A.C. Neuroendocrine targets of endocrine disruptors. Hormones 2010, 9, 16–27. [Google Scholar] [CrossRef] [PubMed]
  240. Sargis, R.M.; Simmons, R.A. Environmental neglect: Endocrine disruptors as underappreciated but potentially modifiable diabetes risk factors. Diabetologia 2019, 62, 1811–1822. [Google Scholar] [CrossRef] [PubMed][Green Version]
  241. Rivollier, F.; Krebs, M.-O.; Kebir, O. Perinatal Exposure to Environmental Endocrine Disruptors in the Emergence of Neurodevelopmental Psychiatric Diseases: A Systematic Review. Int. J. Environ. Res. Public Health 2019, 16, 1318. [Google Scholar] [CrossRef] [PubMed][Green Version]
  242. Basak, S.; Das, M.K.; Duttaroy, A.K. Plastics derived endocrine-disrupting compounds and their effects on early development. Birth Defects Res. 2020, 112, 1308–1325. [Google Scholar] [CrossRef]
  243. Barouki, R. Endocrine disruptors: Revisiting concepts and dogma in toxicology. Comptes Rendus Biol. 2017, 340, 410–413. [Google Scholar] [CrossRef] [PubMed]
  244. Campion, S.; Catlin, N.; Heger, N.; McDonnell, E.V.; Pacheco, S.E.; Saffarini, C.; Sandrof, M.A.; Boekelheide, K. Male Reprotoxicity and Endocrine Disruption. Mol. Clin. Environ. Toxicol. 2012, 101, 315–360. [Google Scholar]
  245. Fénichel, P.; Chevalier, N. Environmental endocrine disruptors: New diabetogens? Comptes Rendus Biol. 2017, 340, 446–452. [Google Scholar] [CrossRef] [PubMed]
  246. Vinggaard, A.M.; Bonefeld-Jørgensen, E.C.; Jensen, T.K.; Fernandez, M.F.; Rosenmai, A.K.; Taxvig, C.; Rodriguez-Carrillo, A.; Wielsøe, M.; Long, M.; Olea, N.; et al. Receptor-based in vitro activities to assess human exposure to chemical mixtures and related health impacts. Environ. Int. 2021, 146, 106191. [Google Scholar] [CrossRef] [PubMed]
  247. Kortenkamp, A. Ten Years of Mixing Cocktails: A Review of Combination Effects of Endocrine-Disrupting Chemicals. Environ. Health Perspect. 2007, 115, 98–105. [Google Scholar] [CrossRef] [PubMed][Green Version]
  248. Petrie, B.; Lopardo, L.; Proctor, K.; Youdan, J.; Barden, R.; Kasprzyk-Hordern, B. Assessment of bisphenol-A in the urban water cycle. Sci. Total. Environ. 2019, 650, 900–907. [Google Scholar] [CrossRef]
Figure 1. Structure of common synthetic and natural endocrine disruptors.
Figure 1. Structure of common synthetic and natural endocrine disruptors.
Water 13 01347 g001
Figure 2. Sources of endocrine disruptors in the aquatic compartment.
Figure 2. Sources of endocrine disruptors in the aquatic compartment.
Water 13 01347 g002
Figure 3. Most frequently detected substances above the environmental quality standard (EQS) at monitoring points in Italy, in 2018 for surface water (a) and groundwater (b).
Figure 3. Most frequently detected substances above the environmental quality standard (EQS) at monitoring points in Italy, in 2018 for surface water (a) and groundwater (b).
Water 13 01347 g003
Figure 4. Schematic representation of the main EDC modes of action: (a) bind to hormone receptors acting as agonist or antagonist; (b) alter signal transduction; (c) block hormone transport across cell membranes; (d) alter hormone synthesis or receptor expression; (e) induce epigenetic modifications.
Figure 4. Schematic representation of the main EDC modes of action: (a) bind to hormone receptors acting as agonist or antagonist; (b) alter signal transduction; (c) block hormone transport across cell membranes; (d) alter hormone synthesis or receptor expression; (e) induce epigenetic modifications.
Water 13 01347 g004
Figure 5. Representation of main endocrine glands targeted by EDCs.
Figure 5. Representation of main endocrine glands targeted by EDCs.
Water 13 01347 g005
Table 1. Concentrations of main EDCs in different water matrices.
Table 1. Concentrations of main EDCs in different water matrices.
Water MatrixEDC TypeAnalytical MethodConcentration
NaproxenHPLC-MS-MS59,300South Africa[74]
IbuprofenHPLC-MS-MS17,600South Africa[74]
CiprofloxacinHPLC-MS-MS14,331South Africa[74]
HPLC-MS-MS45,550South Africa[74]
E2HPLC-DAD510–45,500 Africa[75]
HPLC-MS-MS15,700South Africa[74]
BPAHPLC-DAD3310–15,700 Africa[75]
9-OH risperidoneHPLC-MS-MS0.4Greece[56]
Total pesticidesGC-MS39.3Vietnam[39]
Trialkyl 0.94–16Korea[86]
PhosphatesGC-MS [86]
Chloroalkyl 4.63–67.0Korea[86]
Triclosan 8–103Taiwan[87]
Abbreviations. HPLC-DAD: high-performance liquid chromatograph coupled to a diode array detector.
Table 2. Comparison of the main analytical methods for EDC detection.
Table 2. Comparison of the main analytical methods for EDC detection.
Analytical TechniquesAdvantagesLimitations
Enzyme-linked immunosorbent assay (ELISA)
Identification and quantification
Selective and reproducible
NP (LOD of 6 ng/L [95]), BPA (LOD range 30–80 ng/L [96]), and hormones with good sensitivity (LOD of 0.2–5 ng/L for E2), accuracy and precision
Use of bioantibodies that are unstable
Time consuming
Liquid chromatography methods
Selective and reproducible
Small sample amounts
Limited sample preparation-Identification of a multi-class EDCs
LOD 5.7 ng/L for BPA, 2.7 ng/L for NP, and 3.3 ng/L for E2
High cost
Require expert analysts
Time consuming
Gas chromatography-mass spectrometry (GC-MS)
Identification of organic pollutants
Quantification of small amounts in mass concentration
Suitable for biological matrices and environmental screening
LODs values: 1.5 ng/L for BPA, 0.3 ng/L for NP, and 0.1 ng/L for E2
Requires an expert operator
Time consuming
Derivatization step for non-volatile compounds and polar molecules
Interferences into the sample
High-resolution gas
chromatography-negative chemical ionization-mass spectrometry
Identification of EDC with quickness, accuracy, and high sensitivity
Identification of compounds with functional groups such as phenolic compounds
Identification of complex chemical components
Suitable for mixtures
LOD value: 0.02 ng/L for BPA, 0.05 ng/L for NP, and 0.1 ng/L for E2
Complex to use and expensive
Derivatization treatment
Time consuming
Table 3. Comparison of removal treatment strategies for endocrine-disrupting chemicals in water and wastewater systems.
Table 3. Comparison of removal treatment strategies for endocrine-disrupting chemicals in water and wastewater systems.
Water Source/
EDC Type
Drinking water and wastewater
Pesticides, triclosan, naproxen, ibuprofen, ketoprofen, trimethoprim, acebutolol, diazepam, diltiazem
Great efficiency
Low operative and maintenance costs
No byproducts
Easy to apply
Low energy consumption
Sorbent regeneration or disposal
Use of non-conventional adsorbents enhances the
Adequate contact time and dosage affect the performance
Low removal of carbamazepine and
Emerging compounds, such as PPCPs, pesticides, BPA, E1, E2, EE2, 17β-estradiol-17-acetate, NP, triclosan
Wide spectrum of activity
Ultrafiltration methodology able to remove a high level of all endocrine disruptors
High cost
Toxic waste byproduct
Concentrates (brine) are primarily discharged to the surface water
The challenges of treatment and discharge of the contaminants accumulated during the process
Post treatments
Water and wastewater
Rstrogenic compounds EE2, E2, 17α-acetate, pentachlorophenol, 4tert-octylphenol, triclosan
no toxic substances
High biodegradation
level to 90%
No byproducts
Low costs
Efficacy related to different enzymatic mechanisms
Incubation time
Pretreatment of sample as initial concentration of pollutant
Water and wastewater
E3, BPA, diethylstilbestrol (DES), E2, and EE2, carbamazepine, hormones, phenolic, pesticide, PPCPs, and pharmaceutical compounds, antibiotics (such as ciprofloxacin, amoxicillin, sulfathiazole, and sulfamethazine), nonylphenol deca-ethoxylate
Wide spectrum of efficiency
Removal up to 80% of EDC compounds
High degree of sensitivity
High costs
Regeneration of active substance
Post-treatment water
Publisher’s Note: MDPI stays neutral with regard to jurisdictional claims in published maps and institutional affiliations.

Share and Cite

MDPI and ACS Style

Pironti, C.; Ricciardi, M.; Proto, A.; Bianco, P.M.; Montano, L.; Motta, O. Endocrine-Disrupting Compounds: An Overview on Their Occurrence in the Aquatic Environment and Human Exposure. Water 2021, 13, 1347.

AMA Style

Pironti C, Ricciardi M, Proto A, Bianco PM, Montano L, Motta O. Endocrine-Disrupting Compounds: An Overview on Their Occurrence in the Aquatic Environment and Human Exposure. Water. 2021; 13(10):1347.

Chicago/Turabian Style

Pironti, Concetta, Maria Ricciardi, Antonio Proto, Pietro Massimiliano Bianco, Luigi Montano, and Oriana Motta. 2021. "Endocrine-Disrupting Compounds: An Overview on Their Occurrence in the Aquatic Environment and Human Exposure" Water 13, no. 10: 1347.

Note that from the first issue of 2016, this journal uses article numbers instead of page numbers. See further details here.

Article Metrics

Back to TopTop