Encyclopedia
Please note this is a comparison between Version 1 by Maria João Rocha and Version 5 by Camila Xu.
Synthetic progestins (PGs) are a large family of hormones used in continuously growing amounts in human and animal contraception and medicinal therapies. Because wastewater treatment plants (WWTPs) are unable to eradicate PGs after excretion, they are discharged into aquatic systems, where they can also be regenerated from conjugated PG metabolites. The PGs were considered of particular interest due to their wide use, activity, and hormonal derivation (from testosterone, progesterone, and spirolactone). It is concluded that PGs had been analysed in WWTPs influents and effluents and, to a lesser extent, in other matrices, including surface waters, where their concentrations range from ng/L to a few µg/L. Because of their high affinity for cell hormone receptors, PGs are endocrine disruptor compounds that may alter the reproductive fitness and development of biota. This review focused on their biological effects in fish, which are the most used aquatic model organisms to qualify the impacts of PGs, highlighting the risks that environmental concentrations pose to their health, fecundity, and fertility.
- drospirenone
- EDCs
- estranes
- gestagens
- gonanes
- norpregnanes
- pregnanes
- risk assessment
1. Introduction
Due to water’s vital importance for life, its availability, quality, and governance have been the subject of intense concern, conflicting interests, and heated debate involving communities, industries, governments, and the media [1]. Nonetheless, past actions and the uncontrolled spread of human activities continue to impact water quality and, more broadly, the vast global aquatic ecosystems [2]. One contemporaneous problem widely recognised as serious for mankind is water pollution, including the increase of the concentrations of compounds defined as micropollutants [3][4].
Due to water’s vital importance for life, its availability, quality, and governance have been the subject of intense concern, conflicting interests, and heated debate involving communities, industries, governments, and the media [1]. Nonetheless, past actions and the uncontrolled spread of human activities continue to impact water quality and, more broadly, the vast global aquatic ecosystems [2]. One contemporaneous problem widely recognised as serious for mankind is water pollution, including the increase of the concentrations of compounds defined as micropollutants [3,4].
Water micropollutants are currently mostly anthropogenic in origin and include natural and synthetic compounds that enter the aquatic compartment at concentrations ranging from ng/L to µg/L [5]. Among these contaminants are endocrine disruptor compounds (EDCs). Many of them are active ingredients in hormonal medicines, such as synthetic progestins (also called gestagens, progestogens, or progestins), being of particular concern because they are massively used and designed to act in extremely low dosages in specific cellular receptors [6][7].
Water micropollutants are currently mostly anthropogenic in origin and include natural and synthetic compounds that enter the aquatic compartment at concentrations ranging from ng/L to µg/L [5]. Among these contaminants are endocrine disruptor compounds (EDCs). Many of them are active ingredients in hormonal medicines, such as synthetic progestins (also called gestagens, progestogens, or progestins), being of particular concern because they are massively used and designed to act in extremely low dosages in specific cellular receptors [6,7].
In humans, progestins (PGs) are used instead of progesterone in endocrine therapy due to the rapid metabolisation of the latter hormone [8]. These substances are used not only as contraceptives, as PGs can inhibit ovulation and the proliferation of the endometrium, but also to treat and prevent endometrial hyperplasia and carcinoma [9][10], to control dysfunctional uterine bleeding [11], and even to stimulate the appetite of cancer patients [12]. In veterinary medicine and zootechny, these compounds are also used in therapies of cows and mares (viz. in disorders of the reproductive system) and for estrus synchronisation and preparation of donor and receptor animals in cases of embryo transfer [13].
In humans, progestins (PGs) are used instead of progesterone in endocrine therapy due to the rapid metabolisation of the latter hormone [8]. These substances are used not only as contraceptives, as PGs can inhibit ovulation and the proliferation of the endometrium, but also to treat and prevent endometrial hyperplasia and carcinoma [9,10], to control dysfunctional uterine bleeding [11], and even to stimulate the appetite of cancer patients [12]. In veterinary medicine and zootechny, these compounds are also used in therapies of cows and mares (viz. in disorders of the reproductive system) and for estrus synchronisation and preparation of donor and receptor animals in cases of embryo transfer [13].
2. Classification and Properties of the Most Prominent PGs in Aquatic Environments
PGs are typically classified considering their structural derivation and “generation” (Table 1
). The latter broadly indicates when PGs were introduced to the market. Thus, to understand the effects of PGs, the most relevant classification system is to group them by structure based on the steroid molecule from which they were created; i.e., testosterone, progesterone, and spironolactone [14].
Table 1. Pharmacological groups of the selected progestins referred to in this research considering their structural derivation, generation, and androgenic effects in humans: (+++) highly androgenic; (++) medium androgenic; (+) low androgenic; (-) no androgenic effects.
Pharmacological groups of the selected progestins referred to in this article considering their structural derivation, generation, and androgenic effects in humans: (+++) highly androgenic; (++) medium androgenic; (+) low androgenic; (-) no androgenic effects.
| Hormone | Family | Common Name (Acronym) & CAS | Structure | & Molecular Formula | Generation | & Activity | ||||||
 | | 19-Nortestosterone | Gonanes (C17) or LNG family | Gonanes (C17) or LNG family |
Gestodene ( | GES | ) 60282-87-4 |
 C21H28O2 | C21H28O2 |
3rd Generation 1986 (+++) |
||
| Levonorgestrel ( | LNG | ) 797-63-7 |
 C21H28O2 | C21H28O2 |
2nd Generation 1966 (+++) |
|||||||
| Norgestrel ( | NET | ) 6533-00-2 |
 C21H28O2 | C21H28O2 |
2nd Generation 1966 (+++) |
|||||||
| Etonogestrel ( | ENG | ) 54048-10-1 |
 C22H28O2 | C22H28O2 |
3rd Generation 1998 (+) |
|||||||
Estranes (C18) or NTD family | Estranes (C18) or NTD family |
Norethisterone ( | NTD | ) 68-22-4 |
 C20H26O2 | C20H26O2 |
1st Generation 1951 (++) |
|||||
| Norethisterone acetate ( | NTDA | ) 51-98-9 |
 C22H28O3 | C22H28O3 |
1st Generation 1951 (++) |
|||||||
| Dienogest ( | DIE | ) 65928-58-7 |
 C22H28O2 | C22H28O2 |
4th Generation 1978 (-) |
|||||||
| Progesterone derivatives |  | | 19-Norprogesterone | Norpregnanes (C20) | Norpregnanes (C20) |
Nomegestrol acetate ( | NOMAC | ) 58652-20-3 |
 C23H30O4 | C23H30O4 |
4th Generation 1986 (-) |
|
 17α-Hydroxyprogesterone | 17α-Hydroxyprogesterone |
Pregnanes (C21) | Pregnanes (C21) |
Medroxyprogesterone ( | MEP | ) 520-85-4 |
 C22H32O3 | C22H32O3 |
1st Generation 1957 (+) |
|||
| Medroxyprogesterone acetate ( | MPA | ) 71-58-9 |
 C24H34O4 | C24H34O4 |
1st Generation 1957 (+) |
|||||||
| Megestrol acetate ( | MGA | ) 595-33-5 |
 C22H30O4 | C22H30O4 |
1st Generation 1963 (-) |
|||||||
| Spironolactone | derivative |  Spironolactone | Spironolactone |
Drospirenone ( | DSP | ) 67392-87-4 |
 C24H30O3 | C24H30O3 |
4th Generation 1976 (-) |
|||
Table 1). The gonanes, such as gestodene (GES), norgestrel (NET), and more specifically, its active stereoisomer levonorgestrel (LNG), have high androgenic effects [17]. In contrast, etonogestrel (ENG), which is the biologically active metabolite of desogestrel, is an agonist of the progesterone receptor (PR), showing low androgenic activity and simultaneous glucocorticoid effects [18].
The estranes, NTD and norethisterone acetate (NTDA), have medium androgenic activity [17]. Dienogest (DIE), classified as a fourth generation progestin, is highly specific for the PR [19] and has no androgenic activity [20]. DIE is usually known as a hybrid progestin, as it has the chemical structure of 19-nortestosterone derivatives but shows antiandrogenic activity characteristics, which are typical of progesterone derivatives [20].
The progesterone derivatives, such as those closely related to 19-norprogesterone, which includes nomegestrol acetate (NOMAC), are called “pure” progestational molecules as they bind almost exclusively to the PR and do not interfere with another steroid receptor [19].
In contrast, those PGs derived from 17-hydroxyprogesterone exhibit varying activities. Thus, medroxyprogesterone acetate (MPA) and its metabolite medroxyprogesterone (MEP) have slight androgenic action and exert glucocorticoid activity when given at high doses [21]. Megestrol acetate (MGA) has 50% fewer glucocorticoid effects than MPA [15]. These PGs also act in specific areas of the hypothalamus as antiandrogenic molecules [22]. This action control male sexual behaviour and urine marking—typical of several animals [22]. Moreover, while designed as a PR agonist, MPA has a high binding affinity for glucocorticoid receptors [23][24].
Usually, the most recent PGs derived from progesterone are progestational PGs without androgenic, estrogenic, or glucocorticoid activity. These PGs were conceived to mimic the benefits of progesterone without the undesirable effects of older PGs, such as acne, a decrease in high-density lipoprotein cholesterol (HDL-C), or bloating and water retention [15].
Drospirenone (DSP) is an aldosterone antagonist derived from spironolactone. The primary effect of the latter PG is its anti-mineralocorticoid activity, which causes decreased salt and water retention, leading to lower blood pressure and the absence of androgenic effects [25]. Additionally, DSP exhibits partial antiandrogenic activity [26]—a property that may counter the adverse impact of androgens on hair growth, lipid fluctuation patterns, and insulin, and the possible influence of body composition in postmenopausal women [26]. Further details about PGs’ cellular targets and biological activities in humans can be found in the literature [27][28][29][30].
Presently, the newer formulations of PGs usually contain more potent progestins such as DIE, ENG, and DSP due to their specificity for PR and lack of androgenic effects [30].
3. Waste and Surface Waters Concentrations of Synthetic Progestins
PGs are considered emerging micropollutants in aquatic ecosystems, where they are usually present in concentrations in the order of ng/L. However, accurately knowing their concentrations in waters is crucial since such tiny amounts are potentially harmful to (at least) fish [6][7]. Likely because analysing PGs requires trace analytical methods for their extraction and quantification, the number of studies concerning the environmental levels of these compounds is still scarce and, in a majority, focused on the concentrations of these hormones in influents and effluents from wastewater treatment plants (WWTPs). In addition, the surveyed areas are still limited in space (). The gonanes, such as gestodene (GES), norgestrel (NET), and more specifically, its active stereoisomer levonorgestrel (LNG), have high androgenic effects [17]. In contrast, etonogestrel (ENG), which is the biologically active metabolite of desogestrel, is an agonist of the progesterone receptor (PR), showing low androgenic activity and simultaneous glucocorticoid effects [18].
The estranes, NTD and norethisterone acetate (NTDA), have medium androgenic activity [17]. Dienogest (DIE), classified as a fourth generation progestin, is highly specific for the PR [19] and has no androgenic activity [20]. DIE is usually known as a hybrid progestin, as it has the chemical structure of 19-nortestosterone derivatives but shows antiandrogenic activity characteristics, which are typical of progesterone derivatives [20].
The progesterone derivatives, such as those closely related to 19-norprogesterone, which includes nomegestrol acetate (NOMAC), are called “pure” progestational molecules as they bind almost exclusively to the PR and do not interfere with another steroid receptor [19].
In contrast, those PGs derived from 17-hydroxyprogesterone exhibit varying activities. Thus, medroxyprogesterone acetate (MPA) and its metabolite medroxyprogesterone (MEP) have slight androgenic action and exert glucocorticoid activity when given at high doses [21]. Megestrol acetate (MGA) has 50% fewer glucocorticoid effects than MPA [15]. These PGs also act in specific areas of the hypothalamus as antiandrogenic molecules [22]. This action control male sexual behaviour and urine marking—typical of several animals [22]. Moreover, while designed as a PR agonist, MPA has a high binding affinity for glucocorticoid receptors [23,24].
Usually, the most recent PGs derived from progesterone are progestational PGs without androgenic, estrogenic, or glucocorticoid activity. These PGs were conceived to mimic the benefits of progesterone without the undesirable effects of older PGs, such as acne, a decrease in high-density lipoprotein cholesterol (HDL-C), or bloating and water retention [15].
Drospirenone (DSP) is an aldosterone antagonist derived from spironolactone. The primary effect of the latter PG is its anti-mineralocorticoid activity, which causes decreased salt and water retention, leading to lower blood pressure and the absence of androgenic effects [25]. Additionally, DSP exhibits partial antiandrogenic activity [26]—a property that may counter the adverse impact of androgens on hair growth, lipid fluctuation patterns, and insulin, and the possible influence of body composition in postmenopausal women [26]. Further details about PGs’ cellular targets and biological activities in humans can be found in the literature [27,28,29,30].
Presently, the newer formulations of PGs usually contain more potent progestins such as DIE, ENG, and DSP due to their specificity for PR and lack of androgenic effects [30].
3. Waste and Surface Waters Concentrations of Synthetic Progestins
PGs are considered emerging micropollutants in aquatic ecosystems, where they are usually present in concentrations in the order of ng/L. However, accurately knowing their concentrations in waters is crucial since such tiny amounts are potentially harmful to (at least) fish [6,7]. Likely because analysing PGs requires trace analytical methods for their extraction and quantification, the number of studies concerning the environmental levels of these compounds is still scarce and, in a majority, focused on the concentrations of these hormones in influents and effluents from wastewater treatment plants (WWTPs). In addition, the surveyed areas are still limited in space (Figure 1). From 2015 to 2021, most publications were performed in Europe (48%) and North America (24%). In Asia (19%), South America (5%), and Australia (5%), there are fewer details about the levels of synthetic PGs, and in Africa, as far as researchers notice, there are no data on this subject (
). From 2015 to 2021, most publications were performed in Europe (48%) and North America (24%). In Asia (19%), South America (5%), and Australia (5%), there are fewer details about the levels of synthetic PGs, and in Africa, as far as we notice, there are no data on this subject (
Figure 1
and
Table 2).
).
Figure 1. Locations in which studies on the levels of the synthetic PGs considered in this research were conducted in the aquatic environment from 2015 to 2021 (map generated from
Locations in which studies on the levels of the synthetic PGs considered in this article were conducted in the aquatic environment from 2015 to 2021 (map generated from
https://mapchart.net/world.html, accessed on 27 December 2021).
Besides, there are also differences concerning the types of PGs analysed. For example, in Europe, the most prevalent PGs in Switzerland [31] were DIE and MPA, whilst in the Czech Republic, it was MGA [32], and in Germany [33], it was DIE. In Asia, a recent study showed LNG, DSP, and dydrogesterone as the most frequently detected PGs in China [34].
, accessed on 27 December 2021).
Table 2.
Concentrations of synthetic progestins in waste and surface waters. Average (Av); not detected (ND); not evaluated (n.e.); quantification method (QM); surface waters (S
w
); WWTP influent (WWTPi); WWTP effluents (WWTPe).
| Testosterone derivatives (Gonanes) | PGs | QM | Sw | (ng/L) | WWTPi (ng/L) | WWTPe | (ng/L) | Local (Country) | References | ||
| GES | (1) | 0.2 | 3 | 1 | Basel and canton Zürich WWTPs (Switzerland). | [31] | |||||
| (2) | <0.05 | <0.38–7.7 | <0.29–0.71 | Blanice River and WWTPs (Czech Republic). | [32] | ||||||
| (2) | <0.64 | <0.41–7.0 | <0.19–<3.5 | Several WWTPs (Czech and Slovak Republics) |
[35] | ||||||
| (1) | <0.3 | n.e. | <1.0 | Several WWTPs and rivers (Germany). | [33] | ||||||
| (3) | <0.2 | <3.0 | <1.0 | Jona River and WWTPs (Switzerland). | [36] | ||||||
| (4) | <21.5 | <21.5 | <21.5 | Five WWTPs (Portugal). | [37] | ||||||
| LNG | (1) | <2.5–117 | 493–811 | 32–39 | Langat River Basin (Malaysia). | [38] | |||||
| (5) | <2.5 | n.e. | <2.5 | Southeast Queensland (Australia). | [39] | ||||||
| (6) | 0.85–3.40 | n.e. | n.e. | Lake Balaton (Hungry). | [40] | ||||||
| (7) | <15 | n.e. | <15 | Two WWTPs in Quebec (Canada). | [41] | ||||||
| (2) | <0.08 | <0.26–<2.1 | <0.22–<0.83 | Blanice River and WWTPs (Czech Republic). | [32] | ||||||
| (2) | <0.09 | <0.07–<1.2 | <0.03–<0.32 | Several WWTPs (Czech and Slovak Republics). | [35] | ||||||
| (1) | <0.05–<0.7 | n.e. | <0.3–<1.0 | Several WWTPs and rivers (Germany). | [33] | ||||||
| (1) | ND | ND–38.4 | ND–20.1 | Several WWTPs, Quebec (Canada). | [42] | ||||||
| (8) | <2.5 | <5–299 ± 17 | <3.0 | Québec and Ontario (Canada). | [43] | ||||||
| (4) | n.e. | 2.81 | 1.37 | 21 WWTPs (China). | [34] | ||||||
| (4) | n.e. | n.e. | <1.0 | Several WWTPs effluents (Germany). | [44] | ||||||
| NET | (4) | n.e. | n.e. | <2.0 | Gran Canaria (Spain) | [45] | |||||
| (4) | n.e. | 11.2 | 1.92 | 21 WWTPs (China). | [34] | ||||||
| ENG | (2) | <0.07 | <0.28–<1.4 | <0.21–<0.89 | Blanice River and WWTPs (Czech Republic). | [32] | |||||
| (2) | <0.09 | <0.25–<1.2 | <0.18–<0.94 | Several WWTPs (Czech and Slovak Republics). | [35] | ||||||
| (1) | <0.3 | n.e. | <0.5 | Several WWTPs and rivers (Germany). | [33] | ||||||
| (4) | n.e. | n.e. | <1.2 | Several WWTPs effluents (Germany). | [44] | ||||||
| Testosterone derivatives (Estranes) | NTD | (1) | <2.5–230 | 1048–1137 | 218–265 | Langat River Basin (Malaysia). | [38] | ||||
| (4) | n.e. | n.e. | <2.0 | Gran Canaria (Spain). | [45] | ||||||
| (9) | ND–5.20 | 1.02–94.7 Av. = 25.7 |
ND–1.68 Av. = 1.25 |
Four WWTPs, Shanghai (China). | [46] | ||||||
| (5) | <0.21–3.1 | n.e. | n.e. | Freshwater aquaculture (China). | [47] | ||||||
| (1) | <0.3 | <3 | <0.6 | Basel and canton Zürich WWTPs (Switzerland). | [31] | ||||||
| (7) | <11 | n.e. | <11 | Two WWTPs in Quebec (Canada). | [41] | ||||||
| (2) | <0.04 | <0.02–<0.17 | <0.03–0.85 | Blanice River and WWTPs (Czech Republic). | [32] | ||||||
| (2) | <0.01 | <0.02–<0.91 | <0.02–<4.1 | Several WWTPs (Czech and Slovak Republics). | [35] | ||||||
| (1) | n.e. | n.e. | <0.40 | Pharmaceutical manufacturing facility discharges (USA). | [48] | ||||||
| (3) | <0.3 | <3 | <0.6 | Jona River and several WWTPs (Switzerland). | [36] | ||||||
| (1) | <0.1–<0.3 | n.e. | <1.0 | Several WWTPs and rivers (Germany). | [33] | ||||||
| (8) | 1.7 ± 0.05–2.7 ± 0.17 | <4.8 | 2 ± 0.2–132 ± 2.2 | Québec and Ontario (Canada). | [43] | ||||||
| (10) | <2.3 | <2.3 | <2.3 | Basque Country (Spain). | [49] | ||||||
| (1) | ND | ND–78.8 | ND–31.8 | Several WWTPs, Quebec (Canada). | [42] | ||||||
| (4) | n.e. | 4.02 | 0.20 | 21 WWTPs (China). | [34] | ||||||
| (4) | n.e. | n.e. | <1.0 | Several WWTPs effluents (Germany). | [44] | ||||||
| NTDA | (4) | n.e. | 10.5 | 0.24 | 21 WWTPs (China). | [34] | |||||
| (1) | <0.3 | n.e. | <0.5 | Several WWTPs and rivers (Germany). | [33] | ||||||
| (4) | n.e. | n.e. | <1.0 | Several WWTPs (Germany). | [44] | ||||||
| DIE | (1) | <0.3 | <0.8 | <0.3 | Basel and canton Zürich WWTPs (Switzerland). | [31] | |||||
| (3) | <0.3 | <0.8 | <0.3 | Jona River and several WWTPs (Switzerland). | [36] | ||||||
| (2) | <0.09 | 1.9–11.0 | <0.05–1.0 | Blanice River and WWTPs (Czech Republic). | [32] | ||||||
| (2) | <0.04 | 1.3–12 | <0.04–<4.0 | Several WWTPs (Czech and Slovak Republics). | [35] | ||||||
| (1) | <0.02–2.3 | n.e. | 1.3–4.4 | Several WWTPs and rivers (Germany). | [33] | ||||||
| (4) | n.e. | n.e. | 0.3–3.7 | Several WWTPs effluents (Germany). | [44] | ||||||
| Progesterone derivatives | NOMAC | (2) | <0.07 | <0.08–3.6 | <0.03–0.26 | Blanice River and WWTPs (Czech Republic). | [32] | ||||
| MEP | (5) | <0.07–1.3 | n.e. | n.e. | Freshwater aquaculture (China). | [47] | |||||
| (1) | <0.6 | <6 | <3 | Basel and canton Zürich WWTPs (Switzerland). | [31] | ||||||
| (3) | <0.6 | <6 | <3 | Jona River and several WWTPs (Switzerland). | [36] | ||||||
| (2) | <0.06 | <0.02–<0.13 | <0.03–0.23 | Blanice River and WWTPs (Czech Republic). | [32] | ||||||
| (1) | ND | ND–5.7 | ND–2.9 | Several WWTPs, Quebec (Canada). | [42] | ||||||
| (2) | <0.04 | <0.01–<0.53 | <0.01–0.95 | Several WWTPs (Czech and Slovak Republics). | [35] | ||||||
| (1) | <0.05 | n.e. | <0.08 | Several WWTPs and rivers (Germany). | [33] | ||||||
| MPA | (5) | <0.21–0.31 | n.e. | n.e. | Freshwater aquaculture (China). | [47] | |||||
| (1) | <0.1 | <0.8 | <0.2 | Basel and canton Zürich WWTPs (Switzerland). | [31] | ||||||
| (2) | <0.1 | <0.15–4.4 | <0.09–0.58 | Blanice River and WWTPs (Czech Republic). | [32] | ||||||
| (2) | <0.01 | <0.04–8.1 | <0.04–0.38 | Several WWTPs (Czech and Slovak Republics). | [35] | ||||||
| (3) | <0.1 | <0.8–5.3 | <0.2 | Jona River and several WWTPs (Switzerland). | [36] | ||||||
| (1) | <0.05–0.1 | n.e. | <0.08–<0.3 | Several WWTPs and rivers (Germany). | [33] | ||||||
| (4) | n.e. | 3.09 | 0.23 | 21 WWTPs (China). | [34] | ||||||
| (4) | n.e. | n.e. | <0.6 | Several WWTPs effluents (Germany). | [44] | ||||||
| MGA | (4) | n.e. | n.e. | <60 | Gran Canaria (Spain). | [45] | |||||
| (1) | <0.1 | <1 | <0.6 | Basel and canton Zürich WWTPs (Switzerland). | [31] | ||||||
| (2) | <0.01 | 0.52–13.0 | 0.13–1.0 | Several WWTPs (Czech and Slovak Republics). | [35] | ||||||
| (1) | <0.05–<0.2 | n.e. | <0.06–<0.3 | Several WWTPs and rivers (Germany). | [33] | ||||||
| (2) | <0.07 | <0.03–<6.3 | <0.06–0.4 | Blanice River and WWTPs (Czech Republic). | [32] | ||||||
| (7) | <6–<20 | n.e. | n.e. | Water bodies in Santa Maria (Brazil). | [50] | ||||||
| (4) | n.e. | 0.84 | 0.29 | 21 WWTPs (China). | [34] | ||||||
| Spironolactone | derivative | DSP | (6) | 0.26–4.30 | n.e. | n.e. | Lake Balaton (Hungry). | [40] | |||
| (1) | <0.3 | <4 | <1 | Basel and canton Zürich WWTPs (Switzerland). | [31] | ||||||
| (2) | <0.85 | 0.64–0.77 | <0.18–<0.62 | Blanice River and WWTPs (Czech Republic). | [32] | ||||||
| (2) | <0.04 | 0.34–6.7 | <0.07–<0.29 | Several WWTPs (Czech and Slovak Republics). | [35] | ||||||
| (3) | <0.3 | <4 | <1 | Jona River and several WWTPs (Switzerland). | [36] | ||||||
| (1) | <0.3 | n.e. | <0.05 | Several WWTPs and rivers (Germany). | [33] | ||||||
| (4) | n.e. | 0.69 | 0.39 | 21 WWTPs (China). | [34] | ||||||
| (4) | n.e. | n.e. | <0.8 | Several WWTPs effluents (Germany). | [44] |
(1) Liquid chromatography with tandem mass spectrometry detection (LC-MS/MS); (2) liquid chromatography-tandem atmospheric pressure chemical ionization/atmospheric pressure photoionization with hybrid quadrupole/orbital trap mass spectrometry operated in high-resolution product scan mode (LC-APCI/APPI-HRPS); (3) high-performance liquid chromatography coupled to a triple quadrupole mass spectrometry (HPLC-MS/MS); (4) ultra-performance liquid chromatography coupled with tandem mass detection (UPLC-MS/MS); (5) gas chromatography with tandem mass spectrometry detection (GC-MS/MS); (6) high-performance liquid chromatography-mass spectrometry (HPLC-MS); (7) liquid chromatography–mass spectrometry (LC-MS); (8) triple quadrupole-linear ion trap mass spectrometer using the sMRM (scheduled multiple reaction monitoring) mode (TripleQuad-LIT-MS); (9) rapid resolution liquid chromatography/tandem mass spectrometry (RRLC-MS/MS); (10) laser diode thermal desorption–tandem mass spectrometry (LDTD–MS/MS).
Here, concerning the 12 PGs in
Table 1
, the most investigated (%) were NTD (20%) and LNG (14%). There are still less data concerning MPA, DSP (10%), MEP, MGA (9%), GES, DIE (8%), ENG (5%), NTDA (4%), NET (3%), and NOMAC (1%) (
Table 2).
Therefore, in an accessible and organised way, this entry compiles the existing data in the bibliography relative to the concentrations of 12 PGs from 2015 to 2021, using the “Web of Science Core Collection” and “PubMed” as primary databases. Thus,
).
Therefore, in an accessible and organised way, this paper compiles the existing data in the bibliography relative to the concentrations of 12 PGs from 2015 to 2021, using the “Web of Science Core Collection” and “PubMed” as primary databases. Thus,
Table 2
presents data on the concentrations of these hormones in surface waters and wastewater treatment plants (WWTPs) worldwide, considering their influents and effluents.
Data in
Table 2 were gathered from investigations conducted in various geographic locations, with varying PG inputs, and analysed according to well-established analytical techniques, despite the varying detection and quantification levels and accuracies. It is important to stress that some of the surveyed areas in Asia [38][46][47] are densely populated, which may explain the high amounts of PGs measured in surface waters. Therefore, the disparities between studies from distinct regions are not surprising, corresponding to a wide range of concentrations even when including the three compartments of surface waters, WWTP influents, and WWTP effluents.
Despite the differences mentioned,
were gathered from investigations conducted in various geographic locations, with varying PG inputs, and analysed according to well-established analytical techniques, despite the varying detection and quantification levels and accuracies. It is important to stress that some of the surveyed areas in Asia [38,46,47] are densely populated, which may explain the high amounts of PGs measured in surface waters. Therefore, the disparities between studies from distinct regions are not surprising, corresponding to a wide range of concentrations even when including the three compartments of surface waters, WWTP influents, and WWTP effluents.
Despite the differences mentioned,
Figure 2 shows that synthetic PGs are still present in surface waters in amounts comparable to those observed in WWTP effluents, which is concerning given that dilution is predicted in surface waters. A similar observation was also noticed in previous studies [6][7]. As a result, one can infer that WWTPs do not effectively remove these compounds and/or that some of them can be regenerated in the aquatic environment by deconjugation phenomena (
shows that synthetic PGs are still present in surface waters in amounts comparable to those observed in WWTP effluents, which is concerning given that dilution is predicted in surface waters. A similar observation was also noticed in previous studies [6,7]. As a result, one can infer that WWTPs do not effectively remove these compounds and/or that some of them can be regenerated in the aquatic environment by deconjugation phenomena (
Figure 3).
).
Figure 2.
Data are expressed in boxplots with the minimum, median, maximum, average (+), and interquartile range Q1–Q3. Dots represent average individual values measured in surface waters (S
w
), WWTP influent (WWTPi) and WWTP effluents (WWTPe) around the world concerning PGs derivates from (
A
) Testosterone (n = 42 S
w
, n = 42 WWTPi, and n = 62 WWTPe), (
B
) Progesterone (n = 23 S
w
, n = 22 WWTPi, and n = 29 WWTPe), (
C
) Spirolactone (n = 7 S
w
, n = 7 WWTPi, and n = 9 WWTPe), (
D
) all PGs as a whole (n = 72 S
w
, n = 71 WWTPi and n = 100 WWTPe), (
Figure 3. Sources and pathways for the occurrence of progestins in the environment. The distributions of PGs were based on Besse and Garric (2009) [51].
Sources and pathways for the occurrence of progestins in the environment. The distributions of PGs were based on Besse and Garric (2009) [51].
Figure 2
A shows that PGs derived from testosterone, besides being evaluated in a higher number of studies, were also the hormones with higher concentrations (up to ≅1 µg/L) in the aquatic environments, where their global load reaches ≅97.0% of all PGs considered in
Table 2
vs. 2.49% and 0.57% for progesterone and spironolactone derivatives.
Table 2
reveals that in surface waters, the concentrations of PGs derived from testosterone were typically higher for LNG (<0.05–117 ng/L) and NTD (<0.01–230 ng/L) than those for GES (<0.05–21.5 ng/L), DIE (<0.02–2.3 ng/L), and NTDA (<0.3 ng/L) ≅ ENG (<0.07–<0.3 ng/L). Data concerning NET in surface waters were not available.
In WWTP influents, the concentrations of LNG (<0.07–811 ng/L ) and NTD (<0.02–1137 ng/L) were consistently higher than those of GES (<0.38–<21.5 ng/L), DIE (<0.8–12.0 ng/L) ≅ NET (11.2 ng/L) ≅ NTDA (10.5 ng/L), and ENG (<0.28–<1.4 ng/L).
In WWTP effluents, the highest concentrations were measured for LNG (<0.03–39 ng/L), NTD (<0.03–265 ng/L), followed by GES (<0.19–<21.5 ng/L), DIE (<0.04–4.4 ng/L), NET (<2.0–1.92 ng/L), NTDA (0.24–<1.0 ng/L) ≅ ENG (<0.21–<1.2 ng/L).
Progesterone-derived PGs more commonly exist in surface waters in concentrations ca. 17-fold lower (
Figure 2B) than those reported above for the testosterone derivatives. Such PGs showed similar concentrations to those of the natural hormone progesterone, which ranged from ND to 13.67 ng/L [38][40] in surface waters and from <0.04 ng/L to 24.8 ng/L [32][42] in WWTP influents. In WWTPs effluents, the levels of those PGs were lower than those of progesterone (ND to 110 ng/L) [32][43]. Despite this, progesterone-derived PGs concentrations in surface waters are comparable to those in WWTP effluents, much as testosterone-derived PGs. Moreover, the two most prevalent progesterone-derived PGs, MGA and MEP, were found in identical amounts in all three aquatic compartments.
) than those reported above for the testosterone derivatives. Such PGs showed similar concentrations to those of the natural hormone progesterone, which ranged from ND to 13.67 ng/L [38,40] in surface waters and from <0.04 ng/L to 24.8 ng/L [32,42] in WWTP influents. In WWTPs effluents, the levels of those PGs were lower than those of progesterone (ND to 110 ng/L) [32,43]. Despite this, progesterone-derived PGs concentrations in surface waters are comparable to those in WWTP effluents, much as testosterone-derived PGs. Moreover, the two most prevalent progesterone-derived PGs, MGA and MEP, were found in identical amounts in all three aquatic compartments.
