Phytoestrogens and mycoestrogens in surface waters--Their sources, occurrence, and potential contribution to estrogenic activity.

Environment International 81 (2015) 26–44

Contents lists available at ScienceDirect

Environment International journal homepage: www.elsevier.com/locate/envint

Review

Phytoestrogens and mycoestrogens in surface waters — Their sources, occurrence, and potential contribution to estrogenic activity Barbora Jarošová, Jakub Javůrek, Ondřej Adamovský, Klára Hilscherová ⁎ Research Centre for Toxic Compounds in the Environment (RECETOX), Masaryk University, Kamenice 3, CZ-62500 Brno, Czech Republic

a r t i c l e

i n f o

Article history: Received 7 September 2014 Received in revised form 27 March 2015 Accepted 31 March 2015 Available online xxxx Keywords: Phytoestrogen Mycoestrogen Estrogenicity Sterol Phytosterols Flavonoids Surface waters Estrogenic equivalent Relative potency

a b s t r a c t This review discusses the potential contribution of phytoestrogens and mycoestrogens to in vitro estrogenic activities occurring in surface waters and in vivo estrogenic effects in fish. Main types, sources, and pathways of entry into aquatic environment of these detected compounds were summarized. Reviewed concentrations of phyto/mycoestrogens in surface waters were mostly undetectable or in low ng/L ranges, but exceeded tens of μg/L for the flavonoids biochanin A, daidzein and genistein at some sites. While a few phytosterols were reported to occur at relatively high concentrations in surface waters, information about their potencies in in vitro systems is very limited, and contradictory in some cases. The relative estrogenic activities of compounds (compared to standard estrogen 17β-estradiol) by various in vitro assays were included, and found to differ by orders of magnitude. These potencies were used to estimate total potential estrogenic activities based on chemical analyses of phyto/mycoestrogens. In vivo effective concentrations of waterborne phyto/mycoestrogens were available only for biochanin A, daidzein, formononetin, genistein, equol, sitosterol, and zearalenone. The lowest observable effect concentrations in vivo were reported for the mycoestrogen zearalenone. This compound and especially its metabolites also elicited the highest in vitro estrogenic potencies. Despite the limited information available, the review documents low contribution of phyto/mycoestrogens to estrogenic activity in vast majority of surface waters, but significant contribution to in vitro responses and potentially also to in vivo effects in areas with high concentrations. © 2015 Elsevier Ltd. All rights reserved.

Contents 1. 2.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27 Main types and sources of phyto/mycoestrogens . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27 2.1. Main types of phyto/mycoestrogens . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27 2.2. Primary sources of phyto/mycoestrogens . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27 3. Occurrence and main pathways of phyto/mycoestrogens in surface waters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 28 3.1. Concentrations of phyto/mycoestrogens in surface waters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 28 3.2. Main pathways of phyto/mycoestrogen entry into surface waters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 32 4. Relative potencies (RPs) of phyto/mycoestrogens. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 33 5. Potential contribution of individual phyto/mycoestrogens to total estrogenic activity determined by in vitro assays . . . . . . . . . . . . . . . . . 33 6. Potential contribution of environmental mixtures of phyto/mycoestrogens to total estrogenic activity determined by in vitro assays–concentration addition model . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37 7. In vivo effects of waterborne exposure to phyto/mycoestrogens . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 38 8. Summary of the main identified gaps of knowledge and/or needs for further research . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 41 9. Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42 Acknowledgment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42 Appendix A. Supplementary data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42 References. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42

⁎ Corresponding author. E-mail address: [email protected] (K. Hilscherová).

http://dx.doi.org/10.1016/j.envint.2015.03.019 0160-4120/© 2015 Elsevier Ltd. All rights reserved.

B. Jarošová et al. / Environment International 81 (2015) 26–44

1. Introduction Fish feminization downstream of wastewater treatment plants (WWTPs) as well as some other endocrine-disruptive effects on aquatic organisms have been observed worldwide (WHO and UNEP, 2013). Estrogenic compounds have been shown to play an important role in these effects (Desbrow et al., 1998; Sumpter and Johnson, 2008). In vitro assays evaluating estrogenic activity are widely used nowadays to monitor a variety of environmental waters (e.g., Kinnberg, 2003; Leusch et al., 2010). Compared to in vivo studies and instrumental analyses of estrogenic compounds, in vitro assays are more often used for large-scale monitoring as some have relatively lower cost and higher throughput capacity (Leusch et al., 2010; Liu et al., 2010). The results of in vitro assays reveal the overall estrogenic activity (estrogenicity) of tested samples and are expressed as equivalent concentrations of 17β-estradiol (E2) which would cause the same response as the samples (Estrogenic Equivalents, EEQs). However, the estrogenic potencies of various compounds relative to E2 (i.e., the sensitivity of a particular assay toward a particular estrogen) often differ among bioassays. As a result, the measured values of EEQ in environmental samples can differ greatly among different bioassays (Jarosova et al., 2014; Liu et al., 2010). In vitro estrogenic activity downstream of municipal WWTPs and animal farms can usually be explained by human/animal-excreted steroid estrogens (Sumpter and Johnson, 2008). Alkylphenolic compounds have been identified as major contributors to the estrogenic activity of industrial waste waters (e.g., Johnson et al., 2005). Much less information is available on compounds responsible for estrogenic activity in surface waters at greater distances from WWTP discharges or in headwaters above any WWTP discharges. Moreover, many studies, especially in freshwaters, document that the known man-made xenoestrogens or human/animal-excreted steroid estrogens can explain only a proportion of the measured estrogenic activities (e.g., Fernandez et al., 2007; Sun et al., 2008; Jarosova et al., 2012). Several studies suggested that phyto/mycoestrogens could contribute to the detected activities (Liu et al., 2010). For example, genistein, an isoflavone of plant origin, was identified as the main estrogenic chemical in the Kanzaki River in Japan (Kawanishi et al., 2004). Phytoestrogens can be defined as any plant compounds structurally and/or functionally similar to ovarian and placental estrogens and their active metabolites (Whitten and Patisaul, 2001). They are usually synthesized by plants as metabolites for protection against pathogens and herbivores (Kiparissis et al., 2001), and they also contribute to floral coloration (Clotfelter and Rodriguez, 2006). Phytoestrogens have attracted considerable attention since 1940, when infertility in sheep grazing on pastures rich in subterranean clover (Trifolium subterraneum) in Western Australia was observed (Bennetts et al., 1946). Consequently, a number of phytoestrogens present in plants serving as food for animals and humans have been studied (Murkies et al., 1998; Jefferson et al., 2012). Mycoestrogens are estrogens produced as secondary metabolites of fungi. The well known mycoestrogens are zearalenone (ZEN) and its metabolites primarily produced by the mold Fusarium growing on a variety of crops (e.g., Massart and Saggese, 2010; Metzler et al., 2010). Similarly to phytoestrogens, influence of exposure to mycoestrogens on human/mammalian health via food has been extensively studied (e.g., Massart and Saggese, 2010). In contrast, this study focuses on phyto/mycoestrogens in surface waters and their possible impacts in aquatic environment. Human exposure by this source is negligible compared to exposure by food. The main aim of this paper was to review available information on phyto/mycoestrogens in fresh waters in order to evaluate whether these compounds can substantially contribute to in vitro estrogenic activity and to in vivo effects in surface waters. To address these questions, the relative potencies of a wide range of phytoestrogens and mycoestrogens in a large number of in vitro systems were examined. Furthermore, in vivo effective concentrations of waterborne phyto/

27

mycoestrogens were reviewed. Knowledge on the concentrations of these compounds detected in surface waters was summarized. The potential contribution of phyto/mycoestrogens to the total estrogenic activity found in surface waters was estimated on the basis of the summarized information. 2. Main types and sources of phyto/mycoestrogens 2.1. Main types of phyto/mycoestrogens Most known phytoestrogens belong to the flavonoid chemical group, which is structurally characterized by substituted phenols (at least diphenols) and contains several subclasses (Dixon, 2004). The most well-known subclasses are isoflavones and coumestans (Bacaloni et al., 2005). In the present paper, steroidal flavonoids are classified within another group of phytoestrogens — phytosterols. Lignans comprise another main group of phytoestrogens. Plant lignans are polyphenolic substances derived from phenylalanine (Dixon, 2004). The only so far identified mycoestrogens are ZEN and its metabolites and these are resorcyclic acid lactones (Bucheli et al., 2008; Bakos et al., 2013). These compounds belong within a structurally diverse class of mycotoxins (Murkies et al., 1998). The chemical structures of the most well-known phytoestrogens and mycoestrogens occurring in environmental waters are presented in Fig. 1. In contrast to the abovementioned classification of phytoestrogens, some other authors (e.g., Dixon, 2004; Erbs et al., 2007) define phytoestrogens in a narrower way as only nonsteroidal polyphenols, and, therefore, phytosterols are sometimes considered as a separate group of compounds. 2.2. Primary sources of phyto/mycoestrogens Phytoestrogens occur in numerous plants, including edible species, at rather high concentrations (Liggins et al., 2000; Dixon, 2004). Various types of phytoestrogens can occur in the same plant species and their contents usually differ among different parts of plants (Rochester and Millam, 2009). Here, examples of plants containing high amounts of phytoestrogens are presented. Flavonoids such as genistein, daidzein, and glycitein, and their glycosylated forms, genistin, daidzin, and glycitin, and partially also biochanin A and formononetin have been found at high concentrations in soybean products. Coumestans, biochanin A, and formononetin have been detected mainly in clover (Hoerger et al., 2009; Jefferson et al., 2012). Other plants containing flavonoids are most species of legumes, fruits, and cabbages and some heartwoods of tree species (Murkies et al., 1998; Dixon, 2004; Bacaloni et al., 2005). Phytosterols are plant lipids and therefore mostly occur in plant oils, such as soy, palm, chestnut, and sesame oils (Clotfelter and Rodriguez, 2006). Some phytosterols, particularly sitosterol, have also been found in pulp mill effluents at levels that were shown to cause effects on biota (Mattsson et al., 2001; Leusch and MacLatchy, 2003). Lignans occur in edible plants (i.e., in vegetables such as legumes, fruits, berries, cereals, nuts, and oilseeds) as well as in tree species (Smeds et al., 2007). Their richest food sources are flax seeds and sesame seeds (Smeds et al., 2007). In the intestine, plant lignans are known to be converted by intestinal microbiota to enterolignans, mainly enterodiol and enterolactone (Smeds et al., 2007; Rochester and Millam, 2009). Other sources of phytoestrogens to aquatic ecosystems could be algae, cyanobacteria or aquatic macrophytes forming high biomass in water bodies or some less investigated plants naturally occurring close to surface waters, which contribute plant material into water. However, there is very limited information on the importance of these potential sources. Some marine macroalgae are currently being used as food sources of phytosterols (e.g., Andrade et al., 2013; Kazłowska et al., 2013). Freshwater algae and cyanobacteria have also been shown to contain phytoestrogens (flavonoids and sterols), although information about water concentrations of phytoestrogens produced by these

28


Fig. 1. Chemical structures of phyto/mycoestrogens for which most data on occurrence in surface waters and on relative potencies in in vitro bioassays were available.

organisms is missing (Scholz and Liebezeit, 2012). Studies on phytoestrogens in lakes with cyanobacterial blooms are also missing, in spite of the fact that such blooms are widespread all over the globe and that their frequency is expected to increase further with global warming (Basen et al., 2012). In general, sources of phytoestrogens other than edible plants have been little studied but might also be important. As written above, the mycoestrogen ZEN is produced by fungal species (Fusarium), which are common soil fungi in temperate and warm countries, and are common contaminants of cereal crops worldwide (Bucheli et al., 2008; Massart and Saggese, 2010). Fungi-producing ZEN not only contaminates corn but also colonizes, to a lesser extent, maize, barley, oats, wheat, sorghum, sesame, millet and rice (Massart and Saggese, 2010). Toxin production mainly takes place before harvesting, but may also occur post-harvest (Massart and Saggese, 2010; Waśkiewicz et al., 2012). For example, high concentrations of ZEN metabolites were detected in streams in agricultural areas during spring snowmelt (Kolpin et al., 2014). Fusarium produces ZEN as a major metabolite but it was shown to produce (in much lower concentrations) also α-zearalanol (α-ZAL) and β-zearalanol (β-ZAL) (Erasmuson et al., 1994). After digestion by animals, ZEN is known to be rapidly

metabolized into different metabolites mainly to α-zearalenol (α-ZEL) or β-zearalenol (β-ZEL) and concentration ratios of different metabolites vary among species (Molina-Molina et al., 2014). A-ZEL and βZEL can be either conjugated and excreted or further metabolized to α-ZAL, β-ZAL and zearalanone (ZAN) (Molina-Molina et al., 2014). Moreover, all the five major forms of these estrogens can be further metabolized into all the other forms, although with different efficiencies (Massart and Saggese, 2010). Besides the natural ZEN derivates, α-ZAL has also been synthesized and used as an anabolic growth promoter in beef production in the USA and other countries, whereas anabolic growth promoters are banned in many other countries, including all the member states of the European Union (Erasmuson et al., 1994; Leffers et al., 2001). 3. Occurrence and main pathways of phyto/mycoestrogens in surface waters 3.1. Concentrations of phyto/mycoestrogens in surface waters This section summarizes concentrations of phyto/mycoestrogens reported in published studies, which mostly used LC–MS–MS techniques,


29

Table 1 Detected concentrations of phyto/mycoestrogens in surface waters and their corresponding calculated Estrogenic Equivalents (cEEQ). The cEEQs were calculated by multiplying the concentrations of phyto/mycoestrogens and their relative potencies determined in different studies (as presented in detail in Table 2). Phytoestrogen (group)

Location type

Site (frequency of detection)a, [reference]

Range of concentrations [ng/L]

Mean cEEQsb[ng/L]

Full range of cEEQsc [ng/L]

Biochanin A (flavonoid)

River + coastd River River River River

Douro estuary, Portugal [1] Rivers across Switzerland (38/68) [2] Douro estuary, Portugal [3] Mondego, Portugal [4] Glatt, Töss, and rivers in the Swiss Midlands (212/853), Switzerland [2] Reckenholz (8/8), Switzerland [5] Several rivers (10/23), Switzerland [5] Several rivers, Iowa, USA [6] Toolijooa, Australia [7] Tiber, Italy [8] Minnesota (DS WWTPs), (6/12), USA [9] Tiber, Italy [10] Lake Vadnais (2/12), USA [9] Macquarie Rivulet, Australia [7] Mullet Creek, Australia [7]; [11] Okabena Creek (DS WWTPs), (0/12), USA [9] Straight Lake (0/12), USA [9] Several rivers (1/20), Brazil [12] Tiber, Italy [8] Toolijooa, Australia [7] Macquarie Rivulet, Australia [7] Mullet Creek, Australia [7] Several rivers, Iowa, USA [6] Minnesota (DS WWTPs), (0/12), USA [9] Okabena Creek (DS WWTPs), (0/5), USA [9] Lake Vadnais (0/12), USA [9] Straight Lake (0/12), USA [9] Kanzaki, Japan [13] Zhangcun, China [14] Douro estuary, Portugal [3] Mondego, Portugal [4] Several rivers (2/20), Brazil [12] Toolijooa, Australia [7] Several rivers, Iowa, USA [6] Macquarie Rivulet, Australia [7] Glatt, Töss, and rivers in the Swiss Midlands, (212/853) Switzerland [2] Reckenholz (8/8), Switzerland [5] Douro estuary, Portugal [1] Rivers across Switzerland (4/68) [2] Mullet Creek, Australia [7]; [11] Tiber, Italy [8];[10] Okabena Creek (DS WWTPs), (4/5), USA [9] Minnesota (DS WWTPs), (6/12), USA [9] Lake Vadnais (4/12), USA [9] Rhine, Germany [15] Straight Lake (0/12), USA [9] Tiber, Italy [8] Tiber, Italy [8]

728–19091d nd–297 nd–191 nd–60.2 nd–59.4

0.01–0.38 b0.01 b0.01 b0.01 b0.01

b0.01–2.10 b0.01–0.03 b0.01–0.02 b0.01 b0.01

7–22 nd–12 1.7–5.6 nd–4 1–3 nd–2.3 1–2 nd–1.1 nd–1 nd–0.1 nd nd nd–170 1–2 nd–2 nd–1.5 nd–0.7 nd nd nd nd nd 42,900 nd–1490 nd–884 nd–526 36.2–276 nd–120 10.5–41 14–33 nd–31.5

b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01–0.18 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 1.85 b0.01–0.06 b0.01–0.04 b0.01–0.02 b0.01–0.01 b0.01 b0.01 b0.01 b0.01

b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01–4.59 b0.01–0.05 b0.01–0.05 b0.01–0.04 b0.01–0.02 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01–34.32 b0.01–1.19 b0.01–0.71 b0.01–0.42 b0.01–0.22 b0.01–0.10 b0.01–0.03 b0.01–0.03 b0.01–0.03

5–30 6.7–24.2d nd–17 2–12 2–4 nd–3 nd–1.8 1.6e nd nd nd nd–4

b0.01–0.02 b0.01–0.02 b0.01–0.01 b0.01–0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01

nd–523.8

b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 No relative potency available b0.01–0.20

34–121 nd–67.4 3.2–40.4 nd–22 68–341d nd–242.6 nd–217

0.01–0.05 b0.01–0.03 b0.01–0.02 b0.01 0.01–0.02 b0.01–0.01 b0.01–0.01

b0.01–0.22 b0.01–0.11 b0.01–0.07 b0.01–0.04 b0.01–3.10 b0.01–2.21 b0.01–1.97

44–157 nd–35 5.3–13.5 nd–21 nd–2.4 nd–2 1.1e nd–1 nd nd nd

b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01

b0.01–1.34 b0.01–0.32 b0.01–0.12 b0.01–0.19 b0.01–0.02 b0.01–0.02 b0.01 b0.01 b0.01 b0.01 b0.01

Coumestrol (flavonoid)

Daidzein (flavonoid)

Daidzin (flavonoid) Demethyltexasin (flavonoid) Equol (flavonoid)

Formononetin (flavonoid)

Drainage River River Pond River River River Lake River River River Lake (pristine) River River Pond River River River River River Lake Lake (pristine) River River River River River Pond River River River Drainage River + coastd River River River River River Lake River Lake (pristine) River River River Drainage River River River River + coastd River River Drainage Pond River River River River Lake River River River Lake (pristine)

Glatt, Töss, and rivers in the Swiss Midlands, (212/853) Switzerland [2] Reckenholz (8/8), Switzerland [5] Rivers across Switzerland (21/68) [2] Several rivers, Iowa, USA [6] Several rivers (13/23), Switzerland [5] Douro estuary, Portugal [1] Rivers across Switzerland (60/68) [2] Glatt, Töss, and rivers in the Swiss Midlands, (212/853) Switzerland [2] Reckenholz (8/8), Switzerland [5] Toolijooa, Australia [7] Several rivers, Iowa, USA [6] Several rivers (22/23), Switzerland [5] Minnesota (DS WWTPs), (6/12), USA [9] Macquarie Rivulet, Australia [7] Lake Vadnais (7/12), USA [9] Mullet Creek, Australia [7]; [11] Tiber, Italy [8] Okabena Creek (DS WWTPs), (0/5), USA [9] Straight Lake (0/12), USA [9]

b0.01–0.94

(continued on next page)

30


Table 1 (continued) Phytoestrogen (group)

Location type



Mean cEEQsb[ng/L]


Genistein (flavonoid)

River River River River River River + coastd River River

143,400 nd–2650 nd–507.1 3.96–366 nd–184 17–138 nd–43.8 nd–24.2

34.4 b0.01–0.64 b0.01–0.12 b0.01–0.08 b0.01–0.04 b0.01–0.03 b0.01–0.01 b0.01

0.05–1577 b0.01–29.15 b0.01–5.58 b0.01–4.03 b0.01–2.02 b0.01–1.52 b0.01–0.48 b0.01–0.27

Pond Drainage River River River River River River River River River River River River River Lake River River River River Lake (pristine) River River River

Kanzaki, Japan [13] Zhangcun, China [14] Mondego, Portugal [4] Several rivers (4/20), Brazil [12] Douro estuary, Portugal [3] Douro estuary, Portugal [1] Rivers across Switzerland (6/68) [2] Glatt, Töss, and rivers in the Swiss Midlands, (212/853) Switzerland [2] Toolijooa, Australia [7] Reckenholz, Switzerland [5] Macquarie Rivulet, Australia [7] Several rivers, Iowa, USA [6] Tiber, Italy [10] Fenhe, China [16] Tiber, Italy [8] Siem Reap, Cambodia [16] Ton, Laos [16] Cikamasan, Indonesia [16] Okabena Creek (DS WWTPs), (1/5), USA [9] Minnesota (DS WWTPs), (6/12), USA [9] Long Xuyen city, Vietnam [16] Khong, Thailand [16] Tuaran, Malaysia [16] Lake Vadnais (7/12), USA [9] Mullet Creek, Australia [7] Yeongsan and Seomjin, Korea [16] Khong, Thailand [16] Salut, Malaysia [16] Straight Lake (0/12), USA [9] Kanzaki, Japan [13] Tiber, Italy [8] Tiber, Italy [8]

1–20 nd–14 1–8 nd–8 4–7 3.6–5 3–5 4.4 1.6–3.3 3.2 nd–2.8 nd–2.6 1.5–2.4 nd–1.7 nd–1.7 1.4e nd–1 nd–0.7 nd nd nd nd nd nd–6

b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 No relative potency available

b0.01–0.22 b0.01–0.15 b0.01–0.09 b0.01–0.09 b0.01–0.08 b0.01–0.08 b0.01–0.06 b0.01–0.05 b0.01–0.04 b0.01–0.04 b0.01–0.03 b0.01–0.02 b0.01–0.03 b0.01–0.02 b0.01–0.02 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01 b0.01

Macquarie Rivulet, Australia [7] Toolijooa, Australia [7] Mullet Creek, Australia [7] Possibly cattle affected stream, USA [17] DWTP source water (12/24), USA [18]; [19] Douro estuary, Portugal [1] Sny Magill Creek (control), USA [17] Possibly swine affected stream, USA [17] Rhine, Germany [15] Possibly cattle affected stream, USA [17] DWTP source water (3/24), USA [18] Possibly swine affected stream, USA [17] Sny Magill Creek (control), USA [17] Rhine, Germany [15] Rhine, Germany [15]

0.1–1 nd–0.5 nd–0.5 6000f nd–3000g 9–2528d 2200f 1400f nd 3200f nd–3000 560f 320f 130 nd

Enterodiol (lignan)

Pond

Toolijooa, Australia [7]

0.5–5

1.15 b0.01–0.57 b0.01–0.48 0.42 0.27 b0.01 0.01 b0.01–0.01 b0.01 b0.01 b0.01 No effect (as tested by 1 assay) No effect (as tested by 2 YES assays)

b0.01–4.38 b0.01–2.19 b0.01–1.85 b0.01–1.61 b0.01–1.02 b0.01 b0.01–0.01 b0.01–0.01 b0.01 b0.01 b0.01

Campesterol (sterol)

River Pond River River River River + coastd River River River River River River River River River

River River River Pond Lake River Stream

Macquarie Rivulet, Australia [7] Mullet Creek, Australia [7]; [11] Macquarie Rivulet, Australia [7] Toolijooa, Australia [7] Pilvilampi, Finland [20] Mullet Creek, Australia [7]; [11] 32 streams in agricultural area (26/105), Iowa and Indiana, USA [21] 4 ditches in agricultural areas, Poland [22] Bohdanka and Warta rivers, Poland [23] Several drainages, Switzerland [24] Bohdanka river in urban area, Poland [22] 3 drainages near wheat fields, Poland [23] 15 rivers, Iowa (7/15 in March 2008; 0/15 in June–October 2008), USA [6] Tiber, Italy [10] Gliwice-Sosnice, Poland [25] Rusalka Lake, Poland [23] Minnesota (DS WWTPs), (6/12), USA [9] Gliwice-Sosnice, Poland [25] Odra river, Poland [25] Rhine, Germany [15] Okabena Creek (DS WWTPs), (0/5), USA [9]

0.1–2 nd–0.3 1–74 7–16 12 0.2–5 nd–96

b0.01 b0.01 b0.01 b0.01 b0.10–0.76

b0.01 b0.01 b0.01 b0.01 b0.03–8.05

0.4–80.6 nd–43.7 nd–35 2–25 7.6–12.3 nd–8

b0.01–0.63 b0.01–0.34 b0.01–0.28 0.02–0.20 0.06–0.10 b0.01–0.06

b0.01–6.76 b0.01–3.67 b0.01–2.94 b0.01–2.10 0.02–1.03 b0.01–0.67

2–5 1.52 0.7–1.5 nd–1.2 1.14 0.66 nd nd

0.02–0.04 0.01 b0.01–0.01 b0.01 b0.01 b0.01 b0.08 b0.01

b0.01–0.42 b0.01–0.13 b0.01–0.13 b0.01–0.10 b0.01–0.10 b0.01–0.06 b0.84 b0.08

Genistin (flavonoid) Glycitein (flavonoid)

Sitosterol (sterol)

Stigmasterol (sterol)

Enterolactone (lignan)

Zearalenone (ZEN, mycoestrogen)

Drainage River Drainage River Drainage River River Lake Lake River Melioration ditch River River River


31

Table 1 (continued) Phytoestrogen (group)

α-Zearalanol (α-ZAL, mycoestrogen) β-Zearalanol (α-ZAL, mycoestrogen) α-Zearalenol (α-ZEL, mycoestrogen)

β-Zearalenol (β-ZEL, mycoestrogen)

Location type



Mean cEEQsb[ng/L]


Lake Lake (pristine) River Stream Stream Stream River River River

Lake Vadnais (0/12), USA [9] Straight Lake (0/12), USA [9] Several rivers, Switzerland [24] 2 minimally impacted areas (0/2), USA [21] 3 streams above WWTPs New York (2/3), USA [21] 3 streams below WWTPs New York (1/3), USA [21] Tiber, Italy [10] Rhine, Germany [15] Tiber, Italy [10]

nd nd nd nd Not quantified h Not quantified h nd–3 nd nd–3

b0.01 b0.01 b0.01 b0.10

b0.08 b0.08 b0.13 b1.03

b0.05–0.16 b0.57 b0.02–0.06

b0.02–0.52 b1.73 b0.01–0.14

Stream Stream

3 streams below WWTPs New York (1/3), USA [21] 32 streams in agricultural area (9/105), Iowa and Indiana, USA [21] 3 streams above WWTPs New York (0/3), USA [21] Rhine, Germany [15] 3 streams below WWTPs New York (1/3), USA [21] 32 streams in agricultural area (20/105), Iowa and Indiana, USA [21] 3 streams above WWTPs New York (0/3), USA [21]

nd–817 nd–202

b8.84–388 b8.84–96

b0.46–5719 b0.46–1414

nd nd nd–843 nd–289

b8.84 b4.75 b0.01–1.89 b0.01–0.65

b0.46–b130 b0.25–b70 b0.01–4.22 b0.01–1.45

nd

b0.01

b0.02

Stream River Stream Stream Stream

[1] Rocha et al., 2013; [2] Hoerger et al., 2009; [3] Ribeiro et al., 2009b; [4] Ribeiro et al., 2009a; [5] Erbs et al., 2007; [6] Kolpin et al., 2010; [7] Kang and Price, 2009; [8] Bacaloni et al., 2005; [9] Rearick et al., 2014; [10] Lagana et al., 2004; [11] Kang et al., 2006; [12] Kuster et al., 2009; [13] Kawanishi et al., 2004; [14] Wang et al., 2013; [15] Pawlowski et al., 2003; [16] Duong et al., 2010; [17] Cain et al., 2008; [18] Stackelberg et al., 2007; [19] Stackelberg et al., 2004; [20] Smeds et al., 2007; [21] Kolpin et al., 2014; [22] Waśkiewicz et al., 2012; [23] Gromadzka et al., 2009; [24] Bucheli et al., 2008; [25] Dudziak, 2011. DS WWTPs — downstream of municipal Waste Water Treatment Plants. DWTP — Drinking Water Treatment Plant. nd — not detected or not quantified. a Number of samples (or locations) with detected compound / total number of samples (or locations). b Mean cEEQs were calculated as follows: concentrations of phyto/mycoestrogen were multiplied by geometric mean of relative estrogenic potencies (RPs) of the phyto/mycoestrogen from different in vitro systems (determined in Table 2). c Full ranges of cEEQs were calculated as follows: concentrations of phyto/mycoestrogen were multiplied by min and max RPs of the phyto/mycoestrogen from different in vitro systems (RPs in Table 2). d Concentrations reported as means of 8 locations sampled along the Douro River including 2 locations on the coast. e The study reported means calculated only from positive samples. f Only maximal detected concentrations were reported. g Data taken from graph. h The LOD for zearalenone was 12.3 ng/L, LOQ was not reported in the study.

with a few exceptions using GC–MS, LC–MS, HPLC-DAD or HPLC-UV. The frequencies of detection of phyto/mycoestrogens were generally low (Table 1), especially in studies investigating not only the hot spot locations, which demonstrates they might not be a significant issue in most surface waters. The most studied group of phytoestrogens are flavonoids, especially genistein, daidzein and biochanin A. In most investigated surface waters, concentrations of detected flavonoids occurred in ranges from low ng/L to tens of ng/L (Table 1). At some locations, detected concentrations of flavonoids reached hundreds of ng/L, and, in three studies, concentrations of biochanin A, daidzein and genistein ranged from low μg/L to more than hundred μg/L (Kawanishi et al., 2004; Rocha et al., 2013; Wang et al., 2013). The greatest concentration of biochanin A (19 μg/L) was detected in the estuary of the Douro River in Portugal (Rocha et al., 2013). Apart from this location, biochanin A was detected in Australia, in the USA, and in Europe at concentrations ranging from low ng/L to 60.2 ng/L (Table 1). The reported concentrations of genistein and daidzein ranged from trace levels up to tens of ng/L in Australian, most European, and some Asian studies (Erbs et al., 2007; Kang and Price, 2009; Duong et al., 2010; Kolpin et al., 2010; Hoerger et al., 2009; Rearick et al., 2014). In Brazilian and Portuguese rivers, the concentrations of genistein or daidzein reached hundreds of ng/L (Table 1). Concentrations of genistein and daidzein in samples from one sampling point on the River Zhangcun in China reached 2.65 μg/L and 1.49 μg/L, respectively (Wang et al., 2013). These compounds were detected at even greater levels (143.4 μg/L of genistein and 42.9 μg/L of daidzein) in the Kanzaki River in Japan (Kawanishi et al., 2004). Interestingly, these two studies reporting the greatest concentrations of genistein and daidzein used less specific analytical techniques (HPLC-UV, HPLC-DAD) for their quantification. Two studies attempted to detect the less investigated glycosylated forms of genistein and daidzein – genistin and daidzin – but they were not detected in surface waters (Kawanishi et al., 2004;

Bacaloni et al., 2005). Other flavonoids that were detected at concentrations exceeding 100 ng/L were coumestrol (170 ng/L) in Brazilian rivers (Kuster et al., 2009), equol (121 and 523 ng/L) and formononetin (157 and 217 ng/L) in Swiss rivers and drainage waters (Erbs et al., 2007; Hoerger et al., 2009). Formononetin was also detected in the estuary of the Douro River at concentrations up to 341 ng/L (Rocha et al., 2013). The rest of the investigated flavonoids either occurred in the ng/L range (e.g., Bacaloni et al., 2005; Kang and Price, 2009) or were not detected above detection limits (LODs). Other groups of phytoestrogens have been analyzed in a smaller number of studies. Plant sterols in surface waters were identified in the USA, in Germany, and in Portugal (Pawlowski et al., 2003; Stackelberg et al., 2004, 2007; Cain et al., 2008; Rocha et al., 2013). In the USA, sitosterol and stigmasterol were detected at maximal concentrations of 6000 and 3200 ng/L, respectively, in rivers possibly affected by intense cattle production (Cain et al., 2008). The control river (relatively unaffected by human activity) contained up to 2200 ng/L of sitosterol and 320 ng/L of stigmasterol (Cain et al., 2008). Stigmasterol at a concentration of 130 ng/L but no sitosterol was detected in the River Rhine in Germany (Pawlowski et al., 2003). Concentrations of sitosterol varied from 9 to 2528 ng/L in the Douro River, Portugal (Rocha et al., 2013). The presence of lignans in surface waters was investigated only in one Finnish and two Australian studies (Kang and Price, 2009; Kang et al., 2006; Smeds et al., 2007). Concentrations of enterodiol were from 0.5 to 5 ng/L in a pond and up to 2 ng/L in rivers (Table 1). Concentrations of enterolactone varied from 7 to 16 ng/L in a pond and from 0.2 to 74 ng/L in rivers (Kang and Price, 2009; Kang et al., 2006; Smeds et al., 2007). The occurrence of mycoestrogens has been studied extensively in food and feed products and domestic animals whereas less studies (only US and European) focused on their occurrence in surface waters and most of them analyzed only ZEN (Table 1). Frequencies of detection

32


of ZEN were often relatively low (mostly 0–30%) and differed among sampling seasons (Hartmann et al., 2008; Gromadzka et al., 2009; Maragos, 2012; Waśkiewicz et al., 2012; Kolpin et al., 2014). The highest concentrations of ZEN (up to 96 ng/L) have been detected in small streams or drainage ditches close to fields with crops contaminated by Fusarium sp. (Table 1; Bucheli et al., 2008; Gromadzka et al., 2009; Waśkiewicz et al., 2012; Kolpin et al., 2014). Lower concentrations (from not detected to low ng/L range) were found in rivers and lakes (Pawlowski et al., 2003; Lagana et al., 2004; Bucheli et al., 2008; Gromadzka et al., 2009; Kolpin et al., 2010; Dudziak, 2011; Rearick et al., 2014; Maragos, 2012) with one exception of a Polish river Bogdanka. In studies of the Bogdanka River (Gromadzka et al., 2009; Waśkiewicz et al., 2012), ZEN concentrations were investigated during whole year and the values ranged from LODs (0.3 ng/L) to 44 ng/L depending on sampling season. The highest values were detected during summer and in the early autumn after harvest (Gromadzka et al., 2009; Waśkiewicz et al., 2012). A-ZAL or β-ZAL were investigated in two rivers contaminated by WWTP effluents (Pawlowski et al., 2003; Lagana et al., 2004). The maximal detected concentration of any of these two compounds was 3 ng/L (but the LOD in a study by Pawlowski et al. (2003) was about 10 ng/L). Concentrations of the two other ZEN metabolites α-ZEL and β-ZEL were investigated in Iowa and Indiana agricultural streams and also in streams above and below three WWTPs in New York, USA (Kolpin et al., 2014). The concentrations of α-ZEL and β-ZEL in the agricultural streams were usually below LOD (see the frequencies of detection in Table 1, Kolpin et al., 2014), but the maximal detected concentrations of these compounds repeatedly reached tens or even hundreds of ng/L (Kolpin et al., 2014). No α-ZEL and β-ZEL were detected in waters above the WWTPs in New York but their concentration exceeded 800 ng/L when detected in a stream below one of the WWTPs (Table 1). In contrast, no ZEN, α-ZEL, or ZAN were found in a sample of water from the Rhine River (LODs about 10 ng/L) (Pawlowski et al., 2003). To summarize, the detected levels of mycoestrogens in rivers, streams and drainage ditches usually occurred in the low ng/L or tens of ng/L range. However, much higher concentrations have been reported to occasionally occur, such as the finding of hundreds of ng/L of α-ZEL and β-ZEL in streams in agricultural areas of Iowa or in one site receiving effluent from WWTP in New York (Kolpin et al., 2014). 3.2. Main pathways of phyto/mycoestrogen entry into surface waters The so far assumed main pathways of phytoestrogen entry into the aqueous environment include: a) industrial sources such as leachate and effluents from plant-processing industries, biofuel manufacturing, and paper mills (Stevenson et al., 2011; Wang et al., 2013), b) runoff from agricultural areas, particularly fields cultivated with forage and grain legumes and/or treated with manure (Clotfelter and Rodriguez, 2006; Wang et al., 2013), c) municipal WWTP effluents (Liu et al., 2010), and d) runoff from meadows and orchards in the close surroundings of surface waters (Rocha et al., 2013). Effluents from plant-processing installations including soy-product factories, biofuel manufacturing plants, and paper mills have been studied, for example, by Lundgren and Novak (2009) and by Fernandez et al. (2007). Concentrations of phytoestrogens in these effluents have been reported to be up to tens or even hundreds of μg/L (Lundgren and Novak, 2009); however, the final concentrations of these compounds in surface waters are highly dependent on the hydrological conditions (dilution factors) of the receiving water bodies. Surface waters potentially contaminated by these sources were sampled e.g., by Kawanishi et al. (2004) and by Wang et al. (2013). Kawanishi et al. (2004) found much higher concentrations of genistein and daidzein compared to other studies. The authors described the sampling point as a residential quarter of Osaka City with scattered small and medium-sized industrial factories of many types. There were some possible dischargers including food and wood pulp factories. Quite similar sources of pollution were

identified in the study by Wang et al. (2013), in which samples from the Zhangcun River (Qingdao, China) were collected. The location with the highest concentrations of genistein and daidzein (2.65 and 1.49 μg/L, respectively) received effluents from food factories including manufacturers producing tofu and other bean products (Wang et al., 2013). Runoff from agricultural areas, particularly fields cultivated with forage and grain legumes and/or treated with manure, was investigated in a number of studies. For example, Kolpin et al. (2010) collected samples during the crop growing season from 15 sites across Iowa, USA. Intensive agricultural activity in terms of crops and livestock occurred at all these locations. Also, Erbs et al. (2007) collected drainage samples near fields seeded with Italian ryegrass (Lolium multiflorum) and red clover (Trifolium pratense) in Reckenholz, Switzerland. Low ng/L to hundreds of ng/L of flavonoids were detected in the drainage water as well as in the Iowa rivers. Erbs et al. (2007) and Hoerger et al. (2009) also analyzed samples from many different rivers across Switzerland and the concentrations of detected flavonoids were generally lower compared to the drainage water. Smeds et al. (2007) analyzed humic water of the Pilvilampi Lake in western Finland. The lake was not further described in the study, but in addition to decaying plant material, urine and feces from mammals were mentioned as possible sources of some of the detected lignans (Smeds et al., 2007). In municipal WWTP effluents, phytoestrogens occur due to excretion by people or animals that eat plants (Liu et al., 2010). This source is generally more important in Asian countries, where high amounts of vegetables (especially soya) are consumed, compared to Europe or the USA (Liu et al., 2010). Several studies of phytoestrogens in surface waters mentioned WWTPs located upstream of sampling sites. Among these, Pawlowski et al. (2003), Lagana et al. (2004) and Rearick et al. (2014) detected rather low phytoestrogen concentrations in the Rhine (Germany), Tiber (Italy) and Minnesota (USA) Rivers and also in the Okabena Creek (USA) and the human impacted Lake Vadnais (USA) (Table SI 1 in Supplementary Information). On the other hand, rivers passing through highly populated areas in the state of Rio de Janeiro (although with more sources of pollution than just WWTPs, see Table 3) analyzed in a Brazilian study by Kuster et al. (2009) contained genistein, daidzein and coumestrol at concentrations up to hundreds of ng/L. Stackelberg et al. (2004, 2007), who investigated phytosterols, collected stream-water during low-flow conditions, when the discharge of municipal WWTP effluents accounted for 7–37% of water mass. The authors found sitosterol and stigmasterol in concentrations up to 3000 ng/L (Table 1 and SI 1 in Supplementary Information). Also, the Montego and Douro Rivers (Portugal), which were sampled at their estuaries, passed through several cities with WWTPs. However, in the case of the Montego River, the highest concentrations of phytoestrogens were detected after an extremely dry season, when large forest fires occurred in the upstream catchment (Ribeiro et al., 2009a,b). The phytoestrogens concentrations in the Douro River showed clear seasonal variations indicating that natural sources were probably more important source of these compounds compared to the WWTPs (Rocha et al., 2013). In contrast to the previously mentioned studies, several investigations were aimed at relatively unpolluted locations. Kang et al. (2006) and Kang and Price (2009) focused on Australian rivers and a pond. The Mullet Creek and the Macquarie Rivulet did not contain any WWTP effluents or industrial sources, though some dairy farms or households could have contributed to the presence of phytoestrogens in the water (the detected concentrations are shown in Supplementary Information in Table SI 1). The Macquarie Rivulet, in contrast to Mullet Creek, had a large catchment (Kang et al., 2006; Kang and Price, 2009). The pond contained mainly runoff from a small farm (Kang and Price, 2009). Other less polluted locations were examined in the USA by Cain et al. (2008) and Rearick et al. (2014) and in Italy by Bacaloni et al. (2005). Cain et al. (2008) investigated the presence of phytosterols in a stream relatively unaffected by human activity (Sny Magill Creek) and found concentrations of these compounds comparable to those


found in places potentially affected by concentrated swine and cattle production (Tables 1, and SI 1 in Supplementary Information). None of the five investigated phytoestrogens (flavonoids) was detected in the relatively pristine USA lake (LODs in ng/L range) sampled by Rearick et al. (2014). In the case of the Italian study, water samples were collected from Lake Albano and from the Tiber River. Lake Albano is a basinshaped volcanic depression, which is supplied exclusively by rains and not connected with any wastewater source. None of the nine investigated phytoestrogens was detected in this lake (LODs 0.2–5 ng/L). In contrast, six out of nine investigated phytoestrogens were detected in the River Tiber sampled inside the City of Rome in low ng/L range (Bacaloni et al., 2005, Table SI 1 in Supplementary Information). The identified pathways of mycoestrogens entry into aqueous environment include: a) runoff from fields contaminated by the mold, b) leachate or effluents from storage of the contaminated crops or factories processing them and c) excretion by consumers of the crop (humans via waste waters and animals also via use of manure as a fertilizer) (Metzler et al., 2010; Schwartz et al., 2013). Runoff of ZEN from fields or storage places contaminated by the mold was well investigated in a model emission study (Hartmann et al., 2007, 2008; Bucheli et al., 2008; Schenzel et al., 2012) and was confirmed as an important contamination source in agricultural areas. Also other studies detected ZEN in streams in agricultural areas (Gromadzka et al., 2009; Dudziak, 2011; Waśkiewicz et al., 2012; Kolpin et al., 2014), but it was not always possible to reliably distinguish if its main source was the applied manure, cattle excretion or the crop contaminated by mold (Waśkiewicz et al., 2012; Kolpin et al., 2014). Kolpin et al. (2014) investigated occurrence of mycotoxins including mycoestrogens in intense agricultural areas in the USA and also in 3 WWTPs impacted streams/rivers and both these pathways (agriculture, WWTPs) were concluded to be relevant. At least in some sampling seasons, these authors found higher concentrations of mycotoxins as well as greater maximal level of ZEN, α-ZEL and β-ZEL compared to the previous Swiss study (Hartmann et al., 2007, 2008; Bucheli et al., 2008; Schenzel et al., 2012). The reported possible reason was greater intensity of crop and livestock production. ZEN or some of its metabolites were also detected in several rivers in urban areas including the Polish River Bohdanka (Waśkiewicz et al., 2012), the Tiber River in Italy (Lagana et al., 2004), and some rivers in Minnesota in the USA (Rearick et al., 2014), indicating the relevance of waste waters as a source of these compounds. On the other hand, ZEN was below LOD (b1 ng/L) at some other locations downstream of WWTPs (Rearick et al., 2014) and one study (Dudziak, 2011) demonstrated its high removal of ZEN by WWTP with activated sludge system.

4. Relative potencies (RPs) of phyto/mycoestrogens Various in vitro test systems have been used to assess the RPs of individual phytoestrogens and mycoestrogens: a) transactivation reporter gene (e.g., luciferase, β-galactosidase) assays using mammalian cell lines (MVLN, HEK293T, HGELN, MMV and Ishikawa cells), b) yeastbased reporter gene assays called yeast estrogen screen assays (YES) or recombinant yeast assays (RYA), and c) proliferation tests with estrogen-dependent MCF-7 cells (E-Screen). The relative estrogenic potencies of a wide range of phyto/mycoestrogens determined by the different in vitro systems are summarized in Table 2. The RP of an individual compound is determined as the ratio of the EC50 of standard estrogen E2 and the EC50 of phyto/mycoestrogen. No single study compares the RPs of a wider range of phyto/mycoestrogens from different groups on both yeast and mammalian models. The widest comparison was published by Breinholt and Larsen (1998), who investigated 17 flavonoids by means of E-Screen and YES assays. Similarly, Kalita and Milligan (2010) investigated 14 flavonoids on Ishikawa and YES cells. Bovee et al. (2004) used two YES models with hERα and hERβ to assess the RPs of 6 flavonoids and 2 lignans.

33

Flavonoids have been well investigated, but their RPs varied greatly among studies and the bioassays used. For example, the RP of genistein in YES assays reported by different studies varied by three orders of magnitude (Table 2, Breinholt and Larsen, 1998; Joung et al., 2003). The reason might be the use of different experimental designs and different approaches to data analyses. Since the RPs of phyto/ mycoestrogens can differ widely also among similar types of cells (i.e., yeast, mammalian, or human models), hereinafter we will refer to the RPs of individual assays meaning not only the used model such as YES but also the specific testing protocol used in the particular laboratory (e.g., YES used by Bovee et al. (2004)). In case of most flavonoids, the geometric means of RPs ranged approximately from 10− 3 to 10− 6. The greatest RPs were reported for lesser known flavonoids afrormosin and medicarpin, which however were tested only in one study (Cherdshewasart et al., 2010), and coumestrol, followed by less potent but more abundant equol, genistein, and others (Table 2). There are some contradictory data regarding estrogenic potency of phytosterols. Only a few publications reported RPs of phytosterols and the RPs in case of stigmasterol and campesterol were either relatively low or undetectable (Table 2). Sitosterol showed no response in YES assay (Bovee et al., 2008) and the results from various studies using E-screen are controversial as discussed in detail by Shappell et al. (2012). However, it exhibited consistent estrogenic potency in several tests including two human reporter cell lines (MVLN, HGELN) and E-screen and competitive binding with ERα/ERβ receptors in a study by Gutendorf and Westendorf (2001). Since the goal of our study is to examine if the phyto/mycoestrogens could contribute to the overall estrogenic activity detected by in vitro assays we have used the potencies from responsive assays for further calculations. But it has to be kept in mind that in case of some compounds, especially the sterols, some assays were not responsive and did not show detectable estrogenic activity. There were only two studies that investigated RPs of lignans by in vitro assays and only enterodiol and enterolactone were tested (Bovee et al., 2004; Stanford et al., 2010). Enterolactone showed low estrogenic activity in E-screen assay, and both compounds showed no estrogenicity in YES assay with both alpha and beta estrogen receptors (Table 2). Relatively more information was available for the mycoestrogen ZEN and its metabolites and their RPs were generally higher compared to phytoestrogens (Table 2). In ten out of eleven assays, the RP of ZEN ranged from 2.6 × 10−3 to 8.39 × 10−2. The one outlying assay (Ishikawa cells) showed RP of ZEN as high as 2.59 (Le Guevel and Pakdel, 2001). The authors pointed out low sensitivity of this assay toward 17β-estradiol in their study. Therefore the derived RPs were not included in calculation of geometric mean of RPs and the cEEQ (Section 5). The most potent ZEN metabolite was α-ZEL with the geometric mean of RP of 4.75 × 10−1 whereas β-ZEL was less potent with geometric mean of RP 2.25 × 10−3 (Welshons et al., 1990; Le Guevel and Pakdel, 2001; Minervini et al., 2005; Frizzell et al., 2011). The synthetic compound α-ZAL, which can be also naturally metabolized from ZEN, was also highly potent with the geometric mean of RP 5.2 × 10− 2 and similar RPs were observed also for ZAN and β-ZAL (Table 2). 5. Potential contribution of individual phyto/mycoestrogens to total estrogenic activity determined by in vitro assays The potential contribution of each phyto/mycoestrogen to the total estrogenic activity of surface waters at studied locations determined by in vitro assays was calculated according to Eq. (1): cEEQ ¼ concentration of phyto=mycoestrogen RP

ð1Þ

where cEEQ is the calculated E2-equivalent contributed by a particular phyto/mycoestrogen, and RP is the relative potency of this compound determined in individual bioassay.

34


Table 2 Relative estrogenic potencies of phytoestrogens (relative to 17β-estradiol) as determined by different in vitro systems. Group

Phytoestrogen

Relative potency of E2 (E2 = 1)

Bioassay

Reference

Geometric mean of relative potenciesa

Flavonoids

Afrormosin

1.55 × 10−1 5.27 × 10−2 2.00 × 10−4 1.50 × 10−4 2.40 × 10−6 3.80 × 10−6 No effect 1.10 × 10−4 9.10 × 10−5 5.00 × 10−5 3.10 × 10−5 2.50 × 10−6 1.60 × 10−6 2.96 × 10−8 No effect No effect 2.70 × 10−2 6.70 × 10−3 5.70 × 10−3 1.11 × 10−3 1.00 × 10−3 6.67 × 10−4 6.15 × 10−4 3.33 × 10−4 1.25 × 10−4 1.10 × 10−4 8.00 × 10−4 5.00 × 10−4 2.00 × 10−4 1.10 × 10−4 1.10 × 10−4 3.33 × 10−5 2.71 × 10−5 2.44 × 10−5 1.30 × 10−5 7.89 × 10−6 6.67 × 10−6 2.80 × 10−6 No effect 7.00 × 10−4 2.00 × 10−5 1.60 × 10−6 No effect 2.60 × 10−5 1.80 × 10−3 1.60 × 10−3 1.00 × 10−3 2.30 × 10−4 1.33 × 10−4 1.30 × 10−4 1.00 × 10−4 1.50 × 10−5 9.10 × 10−3 3.00 × 10−4 5.60 × 10−5 7.62 × 10−6 5.37 × 10−6 4.00 × 10−6 1.10 × 10−2 2.00 × 10−3 1.10 × 10−3 1.00 × 10−3 8.00 × 10−4 8.00 × 10−4 5.00 × 10−4 4.90 × 10−4 2.60 × 10−4 2.50 × 10−4 1.50 × 10−4 1.32 × 10−4 1.23 × 10−4 4.50 × 10−5

YES (hERα) YES (hERβ) Ishikawa cells E-Screen YES E-Screen YES E-Screen YES E-Screen E-Screen YES E-Screen E-Screen Ishikawa cells, YES E-Screen, YES YES (hERβ) YES YES (hERα) E-Screen HGELN E-Screen E-Screen MCF7-CAT cells MVLN E-Screen Ishikawa cells E-Screen YES E-Screen YES (hERβ) MCF7-CAT cells E-Screen E-Screen YES E-Screen E-Screen YES YES (hERα) Ishikawa cells YES YES (hERβ) YES (hERα) E-Screen Ishikawa cells YES E-Screen YES E-Screen E-Screen MCF7-CAT cells E-Screen YES (hERα) YES (hERβ) YES E-Screen E-Screen E-Screen YES (hERβ) E-Screen Ishikawa cells YES HGELN E-Screen YES (hERα) YES E-Screen MCF7-CAT cells E-Screen MVLN E-Screen YES

Cherdshewasart et al. (2010)

9.04 × 10−2

Kalita and Milligan (2010) Breinholt and Larsen (1998)

2.29 × 10−5b

Apigenin

Biochanin A

Catechin Caffeic acid Coumestrol

Daidzein

Daidzin

Epicatechin Equol

Eriodictyol Formononetin

Genistein

Stanford et al. (2010) Kalita and Milligan (2010) Breinholt and Larsen (1998) Coldham et al. (1997) Welshons et al. (1990) Stanford et al. (2010) Breinholt and Larsen (1998) Joung et al. (2003) Stanford et al. (2010) Kalita and Milligan (2010) Breinholt and Larsen (1998) Bovee et al. (2004) Coldham et al. (1997) Bovee et al. (2004) Welshons et al. (1990) Gutendorf and Westendorf (2001) Matsumura et al. (2005)c Stanford et al. (2010) Matsumura et al. (2005) Gutendorf and Westendorf (2001) Kalita and Milligan (2010) Matsumura et al. (2005)c Kalita and Milligan (2010) Breinholt and Larsen (1998) Bovee et al. (2004) Matsumura et al. (2005) Stanford et al. (2010) Shappell et al. (2012) Coldham et al. (1997) Joung et al. (2003) Welshons et al. (1990) Breinholt and Larsen (1998) Bovee et al. (2004) Kalita and Milligan (2010)

1.99 × 10−5

1.08 × 10−3

4.31 × 10−5b

2.82 × 10−5b

Bovee et al. (2004) Breinholt and Larsen (1998) Kalita and Milligan (2010)

3.80 × 10−4

Matsumura et al. (2005)c Breinholt and Larsen (1998) Stanford et al. (2010) Breinholt and Larsen (1998) Matsumura et al. (2005) Breinholt and Larsen (1998) Cherdshewasart et al. (2010)

5.41 × 10−5

Coldham et al. (1997) Stanford et al. (2010) Shappell et al. (2012) Welshons et al. (1990) Bovee et al. (2004) Matsumura et al. (2005)c Kalita and Milligan (2010) Kalita and Milligan (2010) Gutendorf and Westendorf (2001) Welshons et al. (1990) Bovee et al. (2004) Coldham et al. (1997) Breinholt and Larsen (1998) Matsumura et al. (2005) Shappell et al. (2012) Gutendorf and Westendorf (2001) Stanford et al. (2010) Breinholt and Larsen (1998)

2.40 × 10−4


35

Table 2 (continued) Group

Phytoestrogen

Genistin

Hesperetin Hexacosaonoic acid 2,3-dihydroxy-propyl ester (palmitic acid glyceryl ester) 7,4′-Dimethoxyisoflavone Chrysin

Isorhamnetin Icaritin

Kaempferol

luteolin Medicarpin Morin Myricetin Naringenin

Phloretin Prunetin Quercetin

Resveratrol

Lignans

Rutin Taxifolin Enterodiol Enterolactone

Phytosterols

Campesterol Sitosterol

Stigmasterol Mycoestrogens

Zearalenone (ZEN)


Bioassay

Reference


1.30 × 10−5 3.42 × 10−7 6.00 × 10−4 1.53 × 10−4 1.00 × 10−4 3.90 × 10−5 b2.00 × 10−5 2.60 × 10−7 no effect 1.00 × 10−4 3.30 × 10−3

E-Screen E-Screen Ishikawa cells E-Screen YES YES (hERβ) YES (hERα) YES Ishikawa cells, YES YES (hERα) YES (hERβ)

Gutendorf and Westendorf (2001) Joung et al. (2003) Kalita and Milligan (2010) Shappell et al. (2012) Kalita and Milligan (2010) Bovee et al. (2004) Breinholt and Larsen (1998) Kalita and Milligan (2010) Cherdshewasart et al. (2010)

5.74 × 10−4

4.00 × 10−3 1.00 × 10−4 2.40 × 10−5 2.00 × 10−6 2.80 × 10−7 1.80 × 10−5 2.30 × 10−7 1.00 × 10−5d 1.00 × 10−5d 1.00 × 10−5d 7.00 × 10−5 1.63 × 10−6 1.10 × 10−6 b2.00 × 10−5 b3.00 × 10−5 5.90 × 10−5 No effect 1.27 × 10−2 6.00 × 10−4 9.60 × 10−6 2.40 × 10−7 2.40 × 10−7 4.00 × 10−4 7.80 × 10−5 7.70 × 10−5 5.20 × 10−5 3.00 × 10−5 1.00 × 10−5 b1.10 × 10−5 2.50 × 10−5 9.40 × 10−7 3.00 × 10−4 2.00 × 10−4 2.00 × 10−6d 1.00 × 10−6d 6.50 × 10−7d 1.60 × 10−7 No effect 2.50 × 10−6 b3.00 × 10−5 b2.00 × 10−5 b2.00 × 10−6 No effect 1.80 × 10−5 No effect No effect 1.07 × 10−6 No effect No effect No effect 7.30 × 10−4 1.00 × 10−4 9.60 × 10−5 No effect No effect 3.30 × 10−6 No effect 2.59e 8.39 × 10−2 1.20 × 10−2f 1.00 × 10−2 9.39 × 10−3

YES (hERβ) YES (hERα) E-Screen E-Screen YES E-Screen YES E-Screen HEK293T (hERα) HEK293T (hERβ) E-Screen E-Screen YES YES Ishikawa cells E-Screen Ishikawa cells, YES YES (hERα) YES (hERβ) E-Screen YES YES Ishikawa cells YES E-Screen YES (hERβ) YES E-Screen YES (hERα) E-Screen YES YES (hERβ) YES (hERα) HEK293T (hERβ) HEK293T (hERα) E-Screen E-Screen Ishikawa cells, YES MCF7-CAT cells Ishikawa cells YES E-Screen Ishikawa cells, YES E-Screen YES (hERβ) YES (hERα) E-Screen YES (hERβ) YES (hERα) YES HGELN cells MVLN E-Screen YES E-Screen YES Ishikawa cells Ishikawa cells RYA (rtER) E-Screen E-Screen MMV


6.32 × 10−4

Breinholt and Larsen (1998) Stanford et al. (2010) Breinholt and Larsen (1998) Breinholt and Larsen (1998)

2.38 × 10−6

Kang et al. (2012)

1.00 × 10−5

Breinholt and Larsen (1998) Stanford et al. (2010) Breinholt and Larsen (1998) Kalita and Milligan (2010)

5.01 × 10−6b

1.38 × 10−4b

2.03 × 10−6

Breinholt and Larsen (1998) Kalita and Milligan (2010) Cherdshewasart et al. (2010)

2.76 × 10−3

Breinholt and Larsen (1998)

1.52 × 10−6

Breinholt and Larsen (1998) Kalita and Milligan (2010) Breinholt and Larsen (1998)

5.78 × 10−5b

Bovee et al. (2004) Kalita and Milligan (2010) Stanford et al. (2010) Bovee et al. (2004) Breinholt and Larsen (1998)

4.85 × 10−6


2.45 × 10−4

Kang et al. (2012)

6.75 × 10−7b

Stanford et al. (2010) Kalita and Milligan (2010) Matsumura et al. (2005) Kalita and Milligan (2010) Matsumura et al. (2005)c Kalita and Milligan (2010) Breinholt and Larsen (1998) Bovee et al. (2004) Stanford et al. (2010) Bovee et al. (2004) Bovee et al. (2008) Gutendorf and Westendorf (2001)

1.91 × 10−4b

Bovee et al. (2008) Shappell et al. (2012) Schmitt et al. (2012) Newill et al. (2007) Le Guevel and Pakdel (2001)

7.87 × 10−3

Minervini et al. (2005) Welshons et al. (1990) Frizzell et al. (2011) (continued on next page)

36


Table 2 (continued) Group

Phytoestrogen

Zearalanone (ZAN)

α-Zearalanol (α-ZAL)

β-Zearalanol (β-ZAL)

α-Zearalenol (α-ZEL)

β-Zearalenol (β-ZEL)


Bioassay

Reference

7.06 × 10−3 5.69 × 10−3 5.00 × 10−3 4.60 × 10−3 4.00 × 10−3 2.60 × 10−3 2.14e 1.63 × 10−1 6.73 × 10−3 5.00 × 10−3 5.00e 1.73 × 10−1 4.40 × 10−2f 1.85 × 10−2 4.69 × 10−1e 4.72 × 10−2 4.50 × 10−2f 4.63 × 10−3 22.73e 7.00f 6.82 × 10−1 4.33 × 10−1 2.47 × 10−2 6.00 × 10−3e 5.0 × 10−3f 3.85 × 10−3 2.64 × 10−3 5.00 × 10−4 b1.04 × 10−4

E-Screen RYA (hER) YES (hERβ) YES (hERα) YES YES Ishikawa cells RYA (rtER) RYA (hER) E-Screen Ishikawa cells RYA (rtER) E-Screen RYA (hER) Ishikawa cells RYA (rtER) E-Screen RYA (hER) Ishikawa cells E-Screen MMV RYA (rtER) RYA (hER) Ishikawa cells E-Screen MMV RYA (hER) E-Screen RYA (rtER)

Shappell et al. (2012) Le Guevel and Pakdel (2001) Bovee et al. (2008)


Schwartz et al. (2010) Coldham et al. (1997) Le Guevel and Pakdel (2001)

1.76 × 10−2

Welshons et al. (1990) Le Guevel and Pakdel (2001)

5.20 × 10−2

Minervini et al. (2005) Le Guevel and Pakdel (2001) Le Guevel and Pakdel (2001)

2.14 × 10−2

Minervini et al. (2005) Le Guevel and Pakdel (2001) Le Guevel and Pakdel (2001) Minervini et al. (2005) Frizzell et al. (2011) Le Guevel and Pakdel (2001) Le Guevel and Pakdel (2001) Minervini et al. (2005) Frizzell et al. (2011) Le Guevel and Pakdel (2001) Welshons et al. (1990) Le Guevel and Pakdel (2001)

4.75 × 10−1

2.25 × 10−3b

rtER: rainbow trout Estrogen Receptor; hER: human Estrogen Receptor; YES: Yeast Estrogen Screen; RYA: Recombinant Yeast Assay. a The geometric means were used to calculate the average cEEQs in Table 1. b Only positive values were included in calculation of geometric mean. c The referred literature showed similar RPs for E-Screen after 7 and 14 days of exposure. RPs after 14 days of exposure were slightly higher and therefore were chosen to be included in this summary Table 2. d Data taken from graph. e The authors pointed out low sensitivity of this assay toward 17β-estradiol in their study. Therefore the derived RPs were not included in calculation of geometric mean of RPs and the cEEQ. f The authors reported RPs as the ratio of EC100 of 17β-estradiol and mycoestrogens (not EC50 as in case of other RPs).

For the most representative characterization of the overall contribution of the individual phyto/mycoestrogens to cEEQ, concentrations (max and min) of each phyto/mycoestrogen detected at each location (or study) were multiplied by the geometric mean of its RPs used as a measure of the average in vitro potency of the compound (Table 1). This mean cEEQs as shown in Tables 1 and 3 (and SI 1 in Supplementary Information) denote the most representative cEEQs across the bioassays. Next to this, the full ranges of potential contribution of these compounds to cEEQ (range of cEEQs from the least sensitive to the most sensitive assays) were derived by multiplying their concentrations (min and max) by their lowest and highest RPs (Table 1). In general, since (as demonstrated in Section 3.1) the investigated phyto/mycoestrogens were frequently not detected at many locations from different studies, their contribution to the overall estrogenic activity at most sites would be negligible. In case of individual flavonoids, estimated cEEQs were also negligible when these compounds occurred at the most commonly found concentrations which were lower than hundreds of ng/L. Only in the case of specifically sensitive YES bioassays would concentrations in tens of ng/L ranges cause a cEEQ of about 0.3 ng/L (Table 1, e.g., formononetin or genistein). When individual flavonoids were present at concentrations of hundreds of ng/L, the cEEQs based on mean RPs reached 0.2 ng/L in case of equol at one site, but were lower for most of the other flavonoids and sites. The cEEQs across assays ranged from negligible (b 0.01) in the case of the least sensitive bioassays to a value of 5.58 ng/L (Table 1, genistein) in the case of the most sensitive assay (YES (hERβ), Bovee et al., 2004, Table 2). There were only a few sites with concentrations exceeding μg/L levels. At one location with extremely high reported concentrations of daidzein and genistein, the mean cEEQ values were 1.85 ng/L and 34.4 ng/L for

daidzein and genistein, respectively (Table 1). The theoretically derived maximal cEEQs in the most sensitive assays at this single location are 34 and 1577 ng/L for daidzein and genistein, respectively. The very high maximal cEEQs for genistein apply only for the case of specifically sensitive YES bioassay (Bovee et al., 2004), while most other assays would show much lower values (represented by the mean cEEQ). In contrast, the cEEQ determined by use of RPs obtained in the least sensitive assay were only b0.01 and 0.05 ng/L for daidzein and genistein, respectively (Table 1). As mentioned in Section 3.1, the greatest reported environmental concentrations of phytosterols seem relatively high but only a few studies determining RPs were available to provide an estimation of cEEQ. In the assays responsive to sitosterol (see discussion in Section 4) cEEQs based on mean RPs were in the range of b0.01 to 1.15 ng/L. The contribution of this compound to cEEQs would reach low ng/L range at the sites with greatest detected concentrations in the most sensitive assays (Tables 1, 3). Lignans did not account for any significant cEEQ (Table 1) because they exert only low or no estrogenicity in the three assays where they were tested (Table 2). Although maximal concentrations of ZEN and its metabolites were lower than those of phytoestrogens, their in vitro potencies were higher and therefore their maximal cEEQs were also relatively high. In most samples, concentrations of these compounds were below LODs (see the frequencies of detection in Table 1), but when detected above the reported LODs (0.4–18.6 ng/L), their cEEQs based on mean RPs ranged from 0.01 up to 388 ng/L. The highest cEEQs were calculated for α-ZEL, which was the most potent estrogen out of all phyto/ mycoestrogens. The cEEQs based on mean RP for zearalenone ranged from b 0.01 to 0.76 ng/L, but it reached up to 8 ng/L in the most sensitive assay (Table 1).


37

Table 3 Individual locations with the highest detected concentrations of phyto/mycoestrogens and their corresponding calculated estrogenic equivalents (cEEQs)a. Group of analyzed compounds, Location type, Reference Flavonoids Drainage water, Reckenholz, Switzerland, Erbs et al. (2007)

Potential sources of the detected phyto/ mycoestrogens suggested in the studies

Phyto/mycoestrogen

Concentrations [ng/L]

Mean cEEQsb [ng/L]


Field seeded with mixture of 56% of Italian ryegrass (Lolium multiflorum) and 43% red clover (T. pratense)

Biochanin A Coumestrol Daidzein Equol Formononetin Genistein Sum of all compounds Biochanin A DAIDZEIN Genistein Sum of all compounds Biochanin A Daidzein Equol Formononetin Genistein Sum of all compounds Coumestrol Daidzein Genistein Sum of all compounds Coumestrol Daidzein Genistein Sum of all compounds Biochanin A Daidzein Genistein Sum of all compounds Daidzein Genistein Sum of all compounds Daidzein Genistein Genistin Sum of all compounds

15 d ndd 15 d 80 d 95d 7d

b0.01 b0.01 b0.01 0.03 b0.01 b0.01 ≥0.03 b0.01 0.03 0.04 ≥0.07 b0.01 b0.01 b0.01 0.01 0.01 ≥0.02 b0.01 0.01 0.08 ≥0.09 0.18 b0.01 b0.01 ≥0.18 b0.01 0.02 0.12 ≥0.14 0.06 0.64 0.70 1.85 34.4 b0.01 36.25

b0.01 b0.01 b0.01–0.01 b0.01–0.13 b0.01–0.86 b0.01–0.08

Douro River at its estuary, Portugal, Ribeiro et al. (2009b)

Vineyards and orchards, indigenous grasses and Fabaceae including red clover, agricultural fields, densely inhabited areas

River Ticiono (a grab sample from a large screening study of rivers across Switzerland) Hoerger et al. (2009)

Probably grassland (for biochanin A and formononetin)

Pavuna Channel near Rio de Janeiro city, Brazil, Kuster et al. (2009)

Households without adequate sanitary infrastructure

River in the Rio de Janeiro city, Brazil, Kuster et al. (2009)

Chemical, pharmaceutical and biotechnology industries, densely inhabited area

Mondego River, Portugal, Ribeiro et al. (2009a)

Corn and rice plantations, indigenous flora of red clover and grasses, forest burning in previous months

River Zhangcun, China, Wang et al. (2013)

Bean product factories (e.g., food factory producing a soya product tofu)

River Kanzaki, Japan, Kawanishi et al. (2004)

Residential quarter of Osaka City with scattered small and medium-sized industrial factories of many types including food and wood pulp factories

nd 597 184 279 nd nd 242.6 43.8 nd 274 336 170 nd 4f 30.9 526 507.1 1490 2650 42,900 143,400 nd

Sterols Stream, Iowa USA, Cain et al. (2008)

Intense cattle production

Sitosterol Stigmasterol Sum of all compounds

6000e 3200e 1.16

Mycoestrogens Stream in agricultural area, Iowa USA, Kolpin et al. (2014)

Extensive agricultural activities such as crop (including soya and corn) and livestock production

Zearalenone α-Zearalenol β-Zearalenol Sum of all compounds Zearalenone α-Zearalenol β-Zearalenol Sum of all compounds

b0.10 96 0.65 97 Detected but not quantified g 817 388 843 1.89 390

Stream below WWTP, New York, USA, Kolpin et al. (2014)

Municipal WWTP with trickling filter technology

nd 202 289

1.15 0.01

b0.01 b0.01–0.48 b0.01–2.02 b0.01–0.03 b0.01 b0.01 b0.01–2.20 b0.01–0.48 b0.01 b0.01–0.22 b0.01–3.70 0.02–4.59 b0.01 b0.01–0.04 b0.01 b0.01–0.42 b0.01–5.58 b0.01–1.19 b0.01–29.15 b0.01–34 0.05–1577 b0.01

b0.01–4.38 b0.01–0.01

b0.03–b1.03 4.99–1414 ≤0.03–1.45

20.2–5719 ≤0.09–4.22

a Individual samples with the highest concentrations of phyto/mycoestrogens from selected studies which reported higher concentrations of phytoestrogens (i.e., more than hundreds of ng/L for flavonoids and/or thousands of ng/L for sterols) or mycoestrogens (i.e., ZEN or its metabolites N 10 ng/L). b Mean cEEQs calculated as described in footnotes to Table 1. c Full ranges of cEEQs calculated as described in footnotes to Table 1. d The concentrations are means of eight samples from the one location sampled from February–March 2007. The frequency of detection was 100% for each of the analyzed phytoestrogens except of coumestrol, which was never detected. e Only maximal detected concentrations were reported. f Data taken from graph. g The LOD for zearalenone was 12.3 ng/L, LOQ was not reported in the study.

6. Potential contribution of environmental mixtures of phyto/ mycoestrogens to total estrogenic activity determined by in vitro assays–concentration addition model Aquatic organisms are not exposed to individual compounds but to complex mixtures. As far as the authors know, there are no studies clarifying interactions (additivity, inhibition, or synergism) among phyto/ mycoestrogens in environmental waters. To derive an estimation of the estrogenic activity which could be induced in in vitro bioassays by environmental mixtures of phyto/mycoestrogens, the concentration

addition (CA, compounds behaving in a clear agonistic manner) model was used. The CA model is generally being used when compounds are expected to act by the same mode of action such as the activation of estrogen receptor. However, some compounds can behave as both an agonist and antagonist of the estrogen receptor if present in different conditions and mixtures (e.g., D'Alessandro et al., 2005) and the effects of mixtures of multiple estrogenic compounds might not be always additive (Alvarez et al., 2013). If the phyto/mycoestrogens behaved in a less than additive manner, the CA model would overestimate the real situation. On the other hand, the actual cEEQ contributions of phyto/

38


mycoestrogens might be greater than indicated by this approach, because just a few selected phyto/mycoestrogens were usually investigated at the locations. Since the concentrations of investigated phyto/mycoestrogens at most locations were below LODs, the in vitro estrogenic activity contributed by analyzed phyto/mycoestrogens would not be of importance in most surface waters. However, at some locations, these compounds can significantly contribute to the in vitro estrogenic activity. Table 3 demonstrates the contribution of several compounds detected at the same locations to cEEQs ordered according to the specific group of phyto/mycoestrogens. There are only a few studies which simultaneously measured phyto/mycoestrogens from more groups (e.g., flavonoids and mycoestrogens in Kolpin et al., 2010 or Lagana et al., 2004), but these studies reported lower levels of the studied compounds (Table SI 1 in Supplementary Information). To demonstrate the situation at most contaminated sites, we identified (if possible) individual samples with the greatest concentrations of studied compounds across studies which reported higher concentrations of phytoestrogens (i.e., more than hundreds of ng/L for flavonoids and/or thousands of ng/L for sterols) and mycoestrogens (i.e., ZEN or its metabolites N 10 ng/L). Consequently, cEEQs based on mean RPs contributed by the different phyto/mycoestrogens detected in these samples were summed (Table 3). The full range of cEEQs across all assays was not summed, since the minimal and maximal RPs for individual compounds frequently originated from different assays. However, the last column in Table 3 includes the information on the greatest cEEQs of individual compounds that would be detected in the most responsive assays. The sums of mean cEEQs contributed by all flavonoids detected in the selected individual samples ranged from about 0.02 to 0.18 ng/L except the two locations with extremely high concentrations of daidzein and genistein. At these locations, the sum of mean cEEQs was about 0.7 ng/L for the River Zhangcun and 36 ng/L for the River Kanzaki (Table 3). In the case of sterols, the sum of mean cEEQs at the location with the highest reported concentrations of sitosterol and stigmatosterol was 1.16 ng/L (maximal cEEQ for sitosterol was 4.38 ng/L). The sums of mean cEEQs for other locations where sterols exceeded thousands of ng/L concentrations (Stackelberg et al., 2004, 2007; Cain et al., 2008; Rocha et al., 2013) would vary from 0.43 to 0.97 ng/L. Two samples with the highest concentrations of mycoestrogens reached sums of mean cEEQs of 97 and 390 ng/L, respectively. Besides the extent of contribution of phyto/mycoestrogens to in vitro estrogenic activities, Table 3 also summarizes the sources (or pathways) of these compounds suggested in listed studies. They include not only agricultural fields, meadows, and livestock production, but also WWTPs, urbanized areas and industries near the rivers. In general, estrogenic activities which cannot be explained only by the presence of vertebrate steroid estrogens, synthetic estrogens and industrial xenoestrogens (e.g., alkylphenols) are being commonly detected in surface waters. The reported “unexplained” estrogenic activities were comparable to the activity estimated based on the highest reported concentrations of phyto/mycoestrogens (Table 3). Several authors reported “unexplained” in vitro estrogenic activity in samples from locations minimally affected by anthropogenic pollution (Danish Ministry of the Environment, 2005; Matthiessen et al., 2006; Nadzialek et al., 2010; Jarosova et al., 2012), which might have been potentially contributed to or even caused by phyto/mycoestrogens. In a Danish national study, almost all samples from background streams showed low (b0.1 ng/L) or no response in YES assay whereas 9 out of 11 samples of reference lakes showed some estrogenic activity with median level of 0.4 ng/L EEQ and a range of 0.1 to 6.2 ng/L EEQ (Danish Ministry of the Environment, 2005). The contributions of the highly potent steroid estrogens ranged only from 0.07 to 0.9 ng/L cEEQ, but the median was similar to the result of the bioassay indicating the “unexplained estrogenic activity” at only few of these locations. Five out of 7 headwaters in relatively unpolluted locations of the Czech Republic contained estrogenic activity when determined by MVLN assay

(Jarosova et al., 2012). EEQ concentrations in water recalculated from concentrations in extract of polar organic integrative samplers ranged from 0.1 to 0.3 ng/L. Low estrogenic activity (0.01–0.03 ng/L EEQ) measured by E-screen was also found at 2 (out of 2) Belgian control rivers (Nadzialek et al., 2010), but neither Nadzialek et al. (2010) nor Jarosova et al. (2012) analyzed the most potent steroid estrogens, which can be excreted not only by humans and livestock but also by wildlife. Matthiessen et al. (2006) investigated estrogenic activity of UK streams up and downstream of livestock farms and, besides others, identified 9 out of 10 relatively pristine upstream samples where the estrogenic activity remained unexplained. The estrogenic activity (as measured by YES and recalculated from concentrations in extract of polar organic integrative samplers) at most of these places ranged from 0.08 to 1.4 ng/L EEQ but it reached 26.5 and 292 ng/L EEQ at two locations (Matthiessen et al., 2006). The authors suggested phytoestrogens as probable contributors to the estrogenic activity and grass silage as their possible candidate source. The contributions of other compounds such as steroid estrogens, estrogenic pesticides, estrogens present in rain water and industrial xenoestrogens were investigated or discussed and concluded to be of very low probability (Matthiessen et al., 2006). In contrast, concentrations of phytoestrogens in μg/L range, which would be needed to cause the 292 ng/L EEQ, were found probable (Matthiessen et al., 2006). Such high concentrations are in a good agreement with some concentrations of phytoestrogens reviewed in this paper (Table 1, Rocha et al., 2013; Kawanishi et al., 2004; Wang et al., 2013). 7. In vivo effects of waterborne exposure to phyto/mycoestrogens Most in vivo studies concerning phyto/mycoestrogens in the aquatic environment focus on dietary sources of phyto/mycoestrogens for fish. This is because the vegetable components of fish food often have high levels (N 30%) of soya and alfalfa — plant sources rich in isoflavonoids (Pelissero and Sumpter, 1992) and can contain cereals contaminated by zearalenone (Bakos et al., 2013). Nevertheless, some studies investigated effects of fish after waterborne exposure to phyto/mycoestrogens (Table 4). These studies focused on the adverse effects of biochanin A, daidzein, equol (a major intestinal metabolite of daidzein), formononetin, genistein, sitosterol, and ZEN. Adverse effects of waterborne biochanin A on juvenile brown trout (Salmo trutta) and larvae as well as on adult zebrafish (Danio rerio) were studied in several experiments by Holbech et al. (2013). The authors exposed brown trout continually for 9 or 10 days and also by 6 h pulse exposure with measurement after 3 days. Plasma vitellogenin (VTG) was increased in all three tests. In the 9 and 10 day exposure tests, the lowest observable effective concentrations (LOEC) were around 10 μg/L, whereas it was 186 μg/L in the 6 h pulse exposure test (Table 4). Unexpected result was obtained when Holbech et al. (2013) performed a long-term (60 days) and a short-term (8 days) exposure of zebrafish fertilized eggs and adults. Increased VTG levels in fish homogenates were observed in the short-term study with LOEC of 114 μg/L but not in the long-term study. Even though no induction of VTG was observed in the long-term study, 209 μg/L of biochanin A caused shift in the sex ratio (more females) in the same test. The authors suggested possible adaptation of fish during the long-term study as a possible explanation but also concluded that further detailed investigation of this phenomenon was needed. In vivo effects of waterborne daidzein and formononetin have only been reported in a single study (Rearick et al., 2014). These authors exposed both adults and larvae of fathead minnow (Pimephales promelas) for 21 days. They observed higher production of eggs by adult fish exposed to 1 μg/L of daidzein and a decrease in the survival of larvae exposed to 1 μg/L of formononetin (Rearick et al., 2014). In the case of both phytoestrogens, as well as when a mixture of these two phytoestrogens and genistein was tested, no effects on the lengths or reaction times of surviving larvae were observed (for details see Table 4).


39

Table 4 Summary of the lowest observable effective concentrations (LOECs) of phytoestrogens obtained in in vivo fish studies. Phytoestrogen

LOEC [μg/L]

Fish species

Age of fish and exposure time

Effect

Reference

Biochanin A

10

Brown trout (Salmo trutta) Brown trout (Salmo trutta) Zebrafish (Danio rerio)

Juveniles, 9 or 10 day exposure (2 similar tests) Juveniles, 6 h pulse exposure, measurement after 3 days 1 day post-fertilization eggs, 60

Induced plasma VTG

Holbech et al. (2013)

Induced plasma VTG


Increased number of females, but not increased VTG in homogenate


Induced VTG in homogenate


No effect on survival, length or reaction times to a threatening stimulus Increased production of eggs but no effects on anatomy, physiology or behavior High incidence of intersex in males (10–87%)

Rearick et al. (2014)

186 209 114 Daidzein

Equol

Formononetin

N1

Adult fish, 21 day exposure

0.4–0.8

Japanese medaka (Oryzias latipes)

1000

Fighting fish (Betta splendens)

Fish soon after hatching exposed until reproductive maturity (=approximately 100 days) Adult males, 28 day exposure

1

Fathead minnow (Pimephales promelas)

New born larvae (N24 h old), 21

Fathead minnow (Pimephales promelas) Japanese medaka (Oryzias latipes)

Adult fish, 21 day exposure

1

N1 10 70 263

270 1–1000

Fathead minnow (Pimephales promelas) Fathead minnow (Pimephales promelas) Japanese medaka (Oryzias latipes) Fathead minnow (Pimephales promelas) Fathead minnow (Pimephales promelas)

Ictalurus punctatus and Sander vitreus Fighting fish (Betta splendens)

day exposure

Fish soon after hatching exposed until reproductive maturity (=approximately 100 days) New born larvae (N24 h old), 21 day exposure Adult fish, 21 day exposure 6-Month old fish, 4-week exposure Juveniles, 21 day exposure Juveniles, 21 day exposure

In vitro exposure of sperm Adult males, 28 day exposure

1000

Japanese medaka (Oryzias latipes)

Fish soon after hatching exposed until reproductive maturity (=approximately 100 days)

N1000


Sexually mature males, 2 day exposure

N1000


2200

Zebrafish (Danio rerio), medaka (Oryzias latipes) Fathead minnow (Pimephales promelas)

Sperm cells removed from mature fish, exposure for short period of sperm activation (4 min) Freshly fertilized embryos, 48 h and 7 day exposure

1 (each chemical)

N1 (each chemical) Sitosterol

day exposure


1

Mixture of genistein, daidzein, and formononetin


1

N1 Genistein

Zebrafish (Danio rerio) Fathead minnow (Pimephales promelas)

day exposure Adult fish, 8 day exposure

75

300 N1



Kiparissis et al. (2003)

Reduced spontaneous swimming activity, reduced aggressive behavior (i.e., delayed responses to the mirror stimulus), increased tendency to build nests Decrease in survival (despite the decrease in survival, larvae that did survive were of similar length and had reaction times to a threatening stimulus comparable to control larvae) No effect on production of eggs, anatomy, physiology or behavior Various gonadal abnormalities (e.g., oocyte atresia increased in females, reduced numbers of oocytes) Decrease in survival

Clotfelter and Rodriguez (2006)

No effect on production of eggs, anatomy, physiology or behavior Induced VTG


Induced VTG in fish homogenates

Panter et al. (2002)

Affected somatic growth (increased weight), length and condition factor also increased but at higher concentration (760 μg/L) Lower ATP content; significantly related to in vitro fertilization rate Reduced aggressive behavior, increased tendency to build nests, but not in a dose–response manner Low incidence (i.e., 12%) of gonadal intersex in males. 72% of males showed feminized secondary sex characteristics No effect on GSI, sperm concentration, motility or fertilization success No effect on sperm motility

Panter et al. (2002)


Rearick et al. (2014) Kiparissis et al. (2003)


Scholz et al. (2004)

Green and Kelly (2008) Clotfelter and Rodriguez (2006) Kiparissis et al. (2003)

Stevenson et al. (2011)

Clotfelter and Gendelman (2014)

Zebrafish embryos showing edema, head and tail deformation

Schiller et al. (2013)

Decrease in survival


No effect on production of eggs, anatomy, physiology or behavior Induction of VTG, no effect on testosterone levels, reduced levels of pregnolon and plasma cholesterol Decreased plasma reproductive steroid levels in males and females No effect on GSI, sperm concentration, motility, or


day exposure Adult fish, 21 day exposure

Fathead minnow (Pimephales promelas) Rainbow trout (Salmo trutta)

Sexually immature fish, 21 day exposure

Gold fish (common varieties) Fighting fish (Betta splendens)

Sexually mature fish, 12 day exposure Sexually mature males, 21 day exposure

Tremblay and Van der Kraak (1999) MacLatchy et al. (1997) Stevenson et al. (2011) (continued on next page)

40


Table 4 (continued) Phytoestrogen

LOEC [μg/L]

Fish species

Age of fish and exposure time

1–1000


Adult males, 28 day exposure

Mixture of phytosterols (80% sitosterol)

10

Lake trout (Salmo trutta lacustris)

Exposure of parent generation 4.5

Mixture of phytosterols (80% sitosterol)

10

Viviparous blenny (Zoarces viviparus)

Parent generation exposed a) from oogenesis to parturition b) from breeding to parturition

Zearalenone

0.002–0.05


Embryos (N24 h post-fertilization), 7 day exposure

0.1–1

Zebrafish (Danio rerio)

months prior to spawning

Spawning groups of adult fish, 21 day pre-exposure, and additional 21 day exposure period

0.32–1


Full life-cycle exposure: 21 day pre-exposure, and 21 day exposure period of F0 adults followed by juvenile exposure up to 140 days and by exposure of F1 adults for 21 days

0.1–5


Embryos (at 1 h post-fertilization), 5 day exposure

0.1–1000


Sexually mature males, 21 day exposure

50–750


Embryos (at 1 h post-fertilization), 5 day exposure

Similarly, adult fish showed no adverse effects on endpoints regarding anatomy, physiology or behavior after exposure to 1 μg/L of these compounds, either exposed individually or in a mixture (Rearick et al., 2014). The effects of genistein and equol have been investigated in more detail because these compounds are the well known to be discharged in effluents from WWTPs (Liu et al., 2010) and from pulp mills (Kiparissis et al., 2001), the adverse effects of which have been observed on fishes (see Introduction). The in vivo studies indicated that the greatest concentrations of phytoestrogens (daidzein and genistein in μg/L or higher ranges) detected in surface waters (Kawanishi et al., 2004; Wang et al., 2013) may be biologically significant. Kiparissis et al. (2003) reported that the waterborne exposure of Japanese medaka (Oryzias latipes) to genistein for 100 days after hatching led to various gonadal abnormalities, some of which were induced by concentrations

Effect fertilization success Nearly no effects on behavior, only an increased tendency to build nests (no dose–response manner evident) Increased dose-dependent egg mortality, smaller egg size, lower mean weight of the yolk sac stage larvae, higher prevalence of deformed or otherwise diseased larvae Higher levels (compared to controls) of phytosterols in larvae and their lower excretion by adult females; stimulated growth of larvae Increased body size (from 0.002 μg/L) and increased frequency of edema (from 0.05 μg/L) but no other deformation. Upregulation of genes for: growth hormone, insulin-like growth factor, luteinizing hormone and vitellogenin, but no upregulation of steroidogenic acute regulatory protein. Decreased spawning frequency and induction of plasma vitellogenin (from 1 μg/L), decreased fecundity (from 0.1 μg/L); no effects on fertility, hatch, survival or gonad morphology. Shifted sex ratio toward female from 0.32 μg/L. Increased wet weight, body length, and condition factor of female fish at 1 μg/L. Induced plasma vitellogenin at 1 μg/L. An increased condition factor in adult female F1 fish implies that exposure of F0 generation to at 1 μg/L zearalenone affected growth of F1 generation. No exposure scenario (0.1–1 μg/L) affected fertility, hatch, embryo survival, and gonad morphology. Induced vitellogenin-1 mRNA expression (observed already from 0.1 μg/L but statistically significant from 5 μg/L) Induction of vitellogenin (protein) (observed from 0.1 μg/L but statistically significant from 1000 μg/L) and induction of expression of vitellogenin-1 mRNA (observed already from 0.1 μg/L but statistically significant from 10 μg/L) Reduced pigmentation, edema and dorsal body axis curvature (LOECs 50, 100 and 750 μg/L). No delay in hatching up to highest tested concentration of 5000 μg/L.

Reference Clotfelter and Rodriguez (2006) Lehtinen et al. (1999)

Mattsson et al. (2001)

Johns et al. (2011)

Schwartz et al. (2010)

Schwartz et al. (2013)

Bakos et al. (2013)

Bakos et al. (2013)

Bakos et al. (2013)

as low as 1 μg/L. Moreover, in the same study, 0.4–0.8 μg/L of equol caused a high incidence of intersex in males (10–87%). A four-weeklong exposure of 6-month-old male Japanese killifish medaka (O. latipes) to 10 μg/L of genistein strongly induced VTG synthesis in vivo (Scholz et al., 2004). One microgram per liter of genistein also caused a decrease in survival of larvae of fathead minnow, although there were no other adverse effects in the surviving larvae (Rearick et al., 2014). In other studies, the determined LOECs of these phytoestrogens were higher (Table 4). For example, Clotfelter and Rodriguez (2006) examined behavioral changes in adult male fighting fish (Betta splendens) as a reaction to 28-day exposure to genistein and equol. Changes in aggressive behavior, an increased tendency to build nests, and, in case of equol, also reduced spontaneous swimming activity were observed at a LOEC of 1000 μg/L (Clotfelter and Rodriguez, 2006). In the same study, some changes were also


observed at lower concentrations of genistein, but did not occur in a dose–response manner. Stevenson et al. (2011) found no effect on gonadosomatic index (GSI), sperm concentration or motility, or fertilization success after a 21-day exposure of fighting fish (B. splendens) to 1 and 1000 μg/L of genistein or a mixture of genistein and sitosterol (1 μg/L each). Similarly, Clotfelter and Gendelman (2014) found no effect on sperm motility in fighting fish when sperm were exposed to 1 and 1000 μg/L of genistein during activation. In contrast, Green and Kelly (2008) incubated testes of channel catfish (Ictalurus punctatus) and walleye (Sander vitreus) in genistein and found a significantly lowered sperm motility time and motility rank, ATP content, and in vitro fertilization rate in both species. The LOEC of the most sensitive endpoint (ATP content) was 270 μg/L of genistein. Panter et al. (2002) tested utility of a 21 day juvenile fathead minnow assay for detection of anti/estrogens including genistein. They found the lowest tested concentration of genistein (70 μg/L) as a LOEC for VTG induction as analyzed in whole body homogenates. Higher concentrations of genistein (263 and 760 μg/L) also increased body weight, length or condition factor at least at some days of conducted analyses (Table 4, Panter et al., 2002). Finally, a new alternative testing approach involving the exposure of fish embryos was performed with zebrafish and medaka for 48 h and 7 days, and the resulting LOEC for genistein was 2200 μg/L, causing edema, and head and tail deformations (Schiller et al., 2013). Species differences in sensitivities can be affected by many factors including previous exposure to phytoestrogens during long-term rearing e.g., from food. Interest in the effects of sitosterol has arisen from studies in which the exposure of fish to bleached pulp mill effluent, in which sitosterol was identified as the dominant plant sterol, resulted in depression in plasma steroid levels, decreased gonadal biosynthetic capacity, and reduced expression of secondary sex characteristics (Munkittrick et al., 1998; Leusch and MacLatchy, 2003). MacLatchy et al. (1997) confirmed that environmentally relevant concentrations occurring in paper mill effluents caused decreases in plasma reproductive steroid levels in goldfish. Lehtinen et al. (1999) and Mattsson et al. (2001) demonstrated adverse effects on larvae of lake trout (S. trutta lacustris) and viviparous blenny (Zoarces viviparus) after exposure of the parent generation to 10 μg/L of mixture of phytosterols containing 80% sitosterol. Several biomarkers such as the induction of VTG were increased in immature rainbow trout after 21-day exposure to 75 μg/L of sitosterol (Tremblay and Van der Kraak, 1999). No effects on gonadosomatic index, sperm concentration, motility, or fertilization success were observed after exposure of adult male fighting fish to 1 μg/L of sitosterol for 21 days (Stevenson et al., 2011). Similarly, only a slightly increased tendency to build nests was observed in a behavioral study investigating changes in aggressive behavior of fighting fish after 96 h exposure to 1–1000 μg/L of sitosterol (Clotfelter and Rodriguez, 2006). In case of ZEN, some adverse in vivo effects have been detected at concentrations comparable to concentrations of ZEN reported in some surface waters (Tables 1 and 4). Johns et al. (2011), who investigated physiological endpoints and expression of genes relevant for growth and reproduction in early life stage of fathead minnow (7 day exposure), found increased body size, edema, and upregulation of most investigated genes at concentrations of ZEN as low as 2–50 ng/L. Other studies investigated effects on embryos, adults or multiple generations of zebrafish (Schwartz et al., 2010, 2013; Bakos et al., 2013). Decreased frequency of spawning, decreased fecundity and induction of plasma VTG were observed after 21 day exposure of adult zebrafish to 0.1– 1 μg/L of ZEN (Table 4, Schwartz et al., 2010). On the other hand, even the highest tested concentration of ZEN (3.2 μg/L) did not affect fertility, hatch, embryo survival or gonad morphology of the tested fish (Schwartz et al., 2010). Similar LOECs (0.32–1 μg/L) were observed in a full life cycle study of zebrafish by Schwartz et al. (2013). In this study, the most sensitive endpoint was sex ratio. Increased number of females occurred already at 0.32 μg/L of ZEN. Other effects (increased wet weight, body length, and condition factor of female fish and

41

induced plasma VTG) occurred at 1 μg/L (Schwartz et al., 2013). Bakos et al. (2013) also exposed embryos (for 5 days) and adults (for 21 days) of zebrafish to ZEN. The authors observed induction of VTG-1 mRNA in embryos and in adults at the same concentration of 0.1 μg/L. However, due to relatively high variability in results the statistically significant LOECs were 5 and 10 μg/L for the embryos and for the adults, respectively. The developmental abnormalities such as reduced pigmentation, edema and dorsal body axis curvature occurred at higher concentrations with LOECs of 50, 100 and 750 μg/L (Bakos et al., 2013). No delay in hatching was observed up to the maximal tested concentration of 5000 μg/L (Bakos et al., 2013). The comparison of the occurrence of phyto/mycoestrogens in surface waters (Table 1 and SI 1 in Supplementary Information) with the available LOEC values (Table 4) demonstrates that these compounds are not present in high concentrations of biological relevance (μg/L) in most sampled locations. On the other hand, the in vivo LOEC was highly exceeded at locations with very high concentrations of genistein reported in the studies by Kawanishi et al. (2004) and Wang et al. (2013). Highest concentrations of equol reported from a Swiss study (Hoerger et al., 2009) exceeded the LOEC for effects in medaka (Kiparissis et al., 2003). Also the study of the Douro River in Portugal (Rocha et al., 2013) revealed concentrations of biochanin A exceeding the reported in vivo LOEC of VTG induction. Except of these studies concentrations of individual flavonoids occurring in surface waters did not exceed the level of μg/L, which approximately represents the lowest known in vivo LOEC for most flavonoids for which in vivo data are available. Sitosterol occurred at concentrations up to several μg/L, relatively close to in vivo LOEC of 10 μg/L for the mixture of phytosterols with 80% sitosterol content (Table 1). However, as previously mentioned, phyto/ mycoestrogens occur in complex mixtures in surface waters and only a few compounds were analyzed in most studies. At some more polluted locations, e.g., the Mondego and the Douro Rivers in Portugal, after the dry hot season in which a large number of forest fires occurred, or during the spring/summer season involving the blooming of meadows and orchards (Ribeiro et al., 2009a,b; Rocha et al., 2013), or urban polluted rivers in Brazil (Kuster et al., 2009), the sum of concentrations of individual phytoestrogens could exceed effective level. Furthermore, many phyto/mycoestrogens including phytosterols were not investigated there. While single phytoestrogens were present below their LOECs, the more likely occurrence of phytoestrogens in mixtures could lead to effects occurring at concentrations under individual LOEC values. Also the concentrations of ZEN reported in drainage waters and some rivers (Table 1) were higher or comparable to LOECs of most in vivo studies. Besides, some of the ZEN metabolites were reported to occur in even higher concentrations and could be even more potent. This indicates that ZEN or its metabolites might contribute to or even pose risk for aquatic biota at contaminated locations. 8. Summary of the main identified gaps of knowledge and/or needs for further research This section summarizes the missing information and shortcomings of literature that have been identified during the analyses of the current state of knowledge regarding potential contribution of phyto/ mycoestrogens to in vitro estrogenic activity and in vivo effects in surface waters. To enable reliable risk assessment of phyto/ mycoestrogens in surface waters and to improve the applicability of in vitro monitoring tools for their detection following knowledge gaps need to be addressed: 1. Only edible plants are commonly investigated as primary sources of phyto/mycoestrogens, although e.g., cyanobacteria, algae and macrophytes abundant in some aquatic environments could be also important source of these compounds. 2. There is no estimation or measurement of the environmental concentrations of some less known flavonoids, which were shown to

42


be relatively potent in vitro. 3. Phytosterols have been repeatedly detected in some surface waters in relatively high concentrations. However, some studies on their in vitro estrogenic potency show contradictory results. More information is needed to show whether these compounds can be monitored by in vitro bioassays and/or which assays would be most suitable for their detection. 4. There is very little information on the occurrence of lignans in surface waters and their in vitro activities or in vivo effects. 5. Only limited information is available regarding concentrations of zearalenone metabolites in aquatic environment. While these compounds showed relatively high in vitro potencies, no study investigated their in vivo effects via water-borne exposure. 6. The relative estrogenic potencies of individual phyto/mycoestrogens can differ among various in vitro assays (even among the same cell types) in orders of magnitude and the bioassays should be well characterized when used for their monitoring. Some less sensitive assays might not be able to detect phyto/mycoestrogens at concentrations that could be relevant for in vivo effects. 7. Always just a few phyto/mycoestrogens have been investigated in the available studies. There is little knowledge about co-occurrence of different groups of phyto/mycoestrogens in surface waters and about interactions of these compounds in environmental mixtures. Even if a single phyto/mycoestrogen is present below its LOEC, the total mixture could cause adverse effects. 8. There are not enough in vivo studies assessing waterborne exposure to phyto/mycoestrogens to conduct reliable risk assessment of these compounds. Moreover, some compounds (e.g., stigmasterol or coumestrol) were reported to occur in surface waters at relatively high concentrations, but no in vivo studies regarding their waterborne exposure are available. 9. Purity and sometimes also the supplier of phyto/mycoestrogens used in in vitro or in vivo studies are often not clearly reported, which complicates the interpretation of the results. Also the potential content of phyto/mycoestrogens in feed provided during the in vivo tests and during previous longer-term rearing of fish is rarely specified. 9. Conclusion Currently available information on phyto/mycoestrogens occurrence in surface waters, their relative in vitro estrogenic potencies, and especially on their in vivo effects after waterborne exposure is limited (Section 8). The review documents that the investigated phyto/mycoestrogens were frequently present at low or undetectable concentrations and thus their contribution to the overall estrogenic activity at vast majority of sites would be negligible. Nevertheless, they can substantially contribute to the estrogenic activity in surface water detected by in vitro assays as well as to in vivo effects at more contaminated sites. Locations near food-processing industries, highly populated areas especially in municipalities with high consumption of vegetables and less developed water treatment systems, waters surrounded by intensive agriculture including pastures, crops, meadows and orchards, and near burning forests have been identified at greatest risk for impact by phytoestrogen exposure. Smaller streams in agricultural areas and below WWTP outflows have been identified at greatest risk for impact by mycoestrogen exposure. Acknowledgment The work was supported by the Czech Science Foundation grant No. GACR P503/12/0553. Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.envint.2015.03.019.

References Alvarez, D.A., Shappell, N.W., Billey, L.O., Bermudez, D.S., Wilson, V.S., Kolpin, D.W., et al., 2013. Bioassay of estrogenicity and chemical analyses of estrogens in streams across the United States associated with livestock operations. Water Res. 47 (10), 3347–3363 (Jun 15). Andrade, P.B., Barbosa, M., Matos, R.P., Lopes, G., Vinholes, J., Mouga, T., et al., 2013. Valuable compounds in macroalgae extracts. Food Chem. 138 (2–3), 1819–1828 (Jun 1). Bacaloni, A., Cavaliere, C., Faberi, A., Foglia, P., Samperi, R., Laganà, A., 2005. Determination of isoflavones and coumestrol in river water and domestic wastewater sewage treatment plants. Anal. Chim. Acta 531 (2), 229–237 (Feb). Bakos, K., Kovács, R., Staszny, Á., Sipos, D.K., Urbányi, B., Müller, F., et al., 2013. Developmental toxicity and estrogenic potency of zearalenone in zebrafish (Danio rerio). Aquat. Toxicol. 15, 136–137 (Jul, 13–21). Basen, T., Rothhaupt, K.O., Martin-Creuzburg, D., 2012. Absence of sterols constrains food quality of cyanobacteria for an invasive freshwater bivalve. Oecologia 170 (1), 57–64 (Sep). Bennetts, H.W., Underwood, E.J., Shier, F.L., 1946. A specific breeding problem of sheep on subterranean clover pastures in Western Australia. Aust. Vet. J. 22, 2–12 (Feb). Bovee, T.F.H., Helsdingen, R.J.R., Rietjens, I.M.C.M., Keijer, J., Hoogenboom, R.L.A.P., 2004. Rapid yeast estrogen bioassays stably expressing human estrogen receptors alpha and beta, and green fluorescent protein: a comparison of different compounds with both receptor types. J. Steroid Biochem. Mol. Biol. 91 (3), 99–109 (Jul). Bovee, T.F.H., Schoonen, W.G.E.J., Hamers, A.R.M., Bento, M.J., Peijnenburg, A.A.C.M., 2008. Screening of synthetic and plant-derived compounds for (anti)estrogenic and (anti)androgenic activities. Anal. Bioanal. Chem. 390 (4), 1111–1119 (Feb). Breinholt, V., Larsen, J.C., 1998. Detection of weak estrogenic flavonoids using a recombinant yeast strain and a modified MCF7 cell proliferation assay. Chem. Res. Toxicol. 622–629. Bucheli, T.D., Wettstein, F.E., Hartmann, N., Erbs, M., Vogelgsang, S., Forrer, H.-R., et al., 2008. Fusarium mycotoxins: overlooked aquatic micropollutants? J. Agric. Food Chem. Am. Chem. Soc. 56 (3), 1029–1034 (Feb 13). Cain, T., Kolpin, D., Vargo, J.D., Wichman, M.D., 2008. Occurrence of Antibiotics, Pharmaceuticals and Sterols at Select Surface and Wastewater Sites in Iowa. pp. 151–157. Cherdshewasart, W., Mahapanichkul, T., Boonchird, C., 2010. Estrogenic and antiestrogenic activities of the Thai traditional herb, Butea superba Roxb. Biosci. Biotechnol. Biochem. 74 (11), 2176–2182. Clotfelter, E.D., Gendelman, H.K., 2014. Exposure to environmentally relevant concentrations of genistein during activation does not affect sperm motility in the fighting fish Betta splendens. Biomed. Res. Int. 2014, 1–5. Clotfelter, E.D., Rodriguez, A.C., 2006. Behavioral changes in fish exposed to phytoestrogens. Environ. Pollut. 144 (3), 833–839 (Dec). Coldham, N., Dave, M., Sivapathasundaram, S., McDonnell, D., Connor, C., Sauerl, M., 1997. Evaluation of a recombinant yeast cell estrogen screening assay. Environ. Health Perspect. 105 (7), 734–742. D'Alessandro, T.L., Boersma-Maland, B.J., Peterson, T.G., Sfakianos, J., Prasain, J.K., Patel, R.P., et al., 2005. Metabolism of phytoestrogen conjugates. Methods Enzymol. 400, 316–342 (Jan). Danish Ministry of the Environment, 2005. Survey of Estrogenic Activity in the Danish Aquatic Environment. p. 170. Desbrow, C., Routledge, E., Brighty, G., Sumpter, J., Waldock, M., 1998. Identification of estrogenic chemicals in STW effluent. 1. Chemical fractionation and in vitro biological screening. Environ. Sci. Technol. 97, 1549–1558. Dixon, R.A., 2004. Phytoestrogens. Annu. Rev. Plant Biol. 55, 225–261 (Jan). Dudziak, M., 2011. Analysis of zearalenone in aqueous environment using GC–MS. Pol. J. Environ. Stud. 20 (1), 231–235 (HARD, POST-OFFICE BOX, 10-718 OLSZTYN 5, POLAND). Duong, C.N., Ra, J.S., Cho, J., Kim, S.D., Choi, H.K., Park, J.-H., et al., 2010. Estrogenic chemicals and estrogenicity in river waters of South Korea and seven Asian countries. Chemosphere 78(3). Elsevier Ltd., pp. 286–293 (Jan). Erasmuson, A.F., Scahill, B.G., West, D.M., 1994. Natural zeranol (a-zearalanol) in the urine of pasture-fed animals. J. Agric. Food Chem. 42, 2721–2725. Erbs, M., Hoerger, C.C., Hartmann, N., Bucheli, T.D., 2007. Quantification of six phytoestrogens at the nanogram per liter level in aqueous environmental samples using 13C3-labeled internal standards. J. Agric. Food Chem. 55 (21), 8339–8345 (Oct 17). Fernandez, M.P., Ikonomou, M.G., Buchanan, I., 2007. An assessment of estrogenic organic contaminants in Canadian wastewaters. Sci. Total Environ. 250–269. Frizzell, C., Ndossi, D., Verhaegen, S., Dahl, E., Eriksen, G., Sørlie, M., et al., 2011. Endocrine disrupting effects of zearalenone, alpha- and beta-zearalenol at the level of nuclear receptor binding and steroidogenesis. Toxicol. Lett. 206 (2), 210–217 (Oct 10). Green, C.C., Kelly, A.M., 2008. Effect of the exogenous soyabean phyto-oestrogen genistein on sperm quality, ATP content and fertilization rates in channel catfish Ictalurus punctatus (Rafinesque) and walleye Sander vitreus (Mitchill). J. Fish Biol. 72 (10), 2485–2499 (Jun). Gromadzka, K., Waśkiewicz, A., Goliński, P., Swietlik, J., 2009. Occurrence of estrogenic mycotoxin — zearalenone in aqueous environmental samples with various NOM content. Water Res. 43 (4), 1051–1059 (Mar). Gutendorf, B., Westendorf, J., 2001. Comparison of an array of in vitro assays for the assessment of the estrogenic potential of natural and synthetic estrogens, phytoestrogens and xenoestrogens. Toxicology 79–89. Hartmann, N., Erbs, M., Wettstein, F.E., Schwarzenbach, R.P., Bucheli, T.D., 2007. Quantification of estrogenic mycotoxins at the ng/L level in aqueous environmental samples using deuterated internal standards. J. Chromatogr. A 1138 (1–2), 132–140 (Jan 5). Hartmann, N., Erbs, M., Wettstein, F.E., Hörger, C.C., Vogelgsang, S., Forrer, H.-R., et al., 2008. Environmental exposure to estrogenic and other myco- and phytotoxins. Chim. Int. J. Chem. 62 (5), 364–367 (May 28).

B. Jarošová et al. / Environment International 81 (2015) 26–44 Hoerger, C.C., Wettstein, F.E., Hungerbühler, K., Bucheli, T.D., 2009. Occurrence and origin of estrogenic isoflavones in Swiss river waters. Environ. Sci. Technol. 43 (16), 6151–6157 (Aug 15). Holbech, H., Schröder, K.D., Nielsen, M.L., Brande-Lavridsen, N., Holbech, B.F., Bjerregaard, P., 2013. Estrogenic effect of the phytoestrogen biochanin A in zebrafish, Danio rerio, and brown trout, Salmo trutta. Aquat. Toxicol. 144–145, 19–25 (Nov 15). Jarosova, B., Blaha, L., Vrana, B., Randak, T., Grabic, R., Giesy, J.P., et al., 2012. Changes in concentrations of hydrophilic organic contaminants and of endocrine-disrupting potential downstream of small communities located adjacent to headwaters. Environ. Int. 45, 22–31 (Sep 15). Jarosova, B., Blaha, L., Giesy, J.P., Hilscherova, K., 2014. What level of estrogenic activity determined by in vitro assays in municipal waste waters can be considered as safe? Environ. Int. 64, 98–109 (Dec 30). Jefferson, W.N., Patisaul, H.B., Williams, C.J., 2012. Reproductive consequences of developmental phytoestrogen exposure. Reproduction 143 (3), 247–260 (Mar). Johns, S.M., Denslow, N.D., Kane, M.D., Watanabe, K.H., Orlando, E.F., Sepúlveda, M.S., 2011. Effects of estrogens and antiestrogens on gene expression of fathead minnow (Pimephales promelas) early life stages. Environ. Toxicol. 26 (2), 195–206 (Apr). Johnson, A.C., Aerni, H.R., Gerritsen, A., Gibert, M., Giger, W., Hylland, K., et al., 2005. Comparing steroid estrogen, and nonylphenol content across a range of European sewage plants with different treatment and management practices. Water Res. 47–58. Joung, K.E., Kim, Y.W., Sheen, Y.Y., 2003. Assessment of the estrogenicity of isoflavonoids, using MCF-7-ERE-Luc cells. Arch. Pharm. Res. 26 (9), 756–762 (Sep). Kalita, J.C., Milligan, S.R., 2010. In vitro estrogenic potency of phytoestrogen-glycosides and some plant flavanoids. Indian J. Sci. Technol. 3 (12), 1142–1147. Kang, J., Price, W.E., 2009. Occurrence of phytoestrogens in municipal wastewater and surface waters. J. Environ. Monit. 11 (Ml), 1477–1483 (Aug). Kang, J., Price, W.E., Hick, L.A., 2006. Simultaneous determination of isoflavones and lignans at trace levels in natural waters and wastewater samples using liquid chromatography/electrospray ionization ion trap mass spectrometry. Rapid Commun. Mass Spectrom. 20 (16), 2411–2418 (Jan). Kang, H.K., Choi, Y.-H., Kwon, H., Lee, S.-B., Kim, D.-H., Sung, C.K., et al., 2012. Estrogenic/ antiestrogenic activities of an Epimedium koreanum extract and its major components: in vitro and in vivo studies. Food Chem. Toxicol. 50 (8), 2751–2759 (Aug). Kawanishi, M., Takamura-Enya, T., Ermawati, R., Shimohara, C., Sakamoto, M., Matsukawa, K., et al., 2004. Detection of genistein as an estrogenic contaminant of river water in Osaka. Environ. Sci. Technol. 6424–6429. Kazłowska, K., Lin, H.-T.V., Chang, S.-H., Tsai, G.-J., 2013. In vitro and in vivo anticancer effects of sterol fraction from red algae Porphyra dentata. Evid. Based Complement. Alternat. Med. 2013, 493869 (Jan). Kinnberg, K., 2003. Evaluation of In Vitro Assays for Determination of Estrogenic Activity in the Environment. Danish Environmental Protection Agency, Danish Ministry of the Environment. Kiparissis, Y., Hughes, R., Metcalfe, C., Ternes, T., 2001. Identification of the isoflavonoid genistein in bleached Kraft mill effluent. Environ. Sci. Technol. 2423–2427. Kiparissis, Y., Balch, G.C., Metcalfe, T.L., Metcalfe, C.D., 2003. Effects of the isoflavones genistein and equol on the gonadal development of Japanese medaka (Oryzias latipes). Environ. Health Perspect. 111 (9), 1158–1163 (Feb 5). Kolpin, D.W., Hoerger, C.C., Meyer, M.T., Wettstein, F.E., Hubbard, L.E., Bucheli, T.D., 2010. Phytoestrogens and mycotoxins in Iowa streams: an examination of underinvestigated compounds in agricultural basins. J. Environ. Qual. 39 (6), 2089–2099. Kolpin, D.W., Schenzel, J., Meyer, M.T., Phillips, P.J., Hubbard, L.E., Scott, T.-M., et al., 2014. Mycotoxins: diffuse and point source contributions of natural contaminants of emerging concern to streams. Sci. Total Environ. 470–471 (Feb 1, 669–76). Kuster, M., Azevedo, D.A., López de Alda, M.J., Aquino Neto, F.R., Barceló, D., 2009. Analysis of phytoestrogens, progestogens and estrogens in environmental waters from Rio de Janeiro (Brazil). Environ Int. 35(7). Elsevier Ltd., pp. 997–1003 (Oct) Lagana, A., Bacaloni, A., De Leva, I., Faberi, A., Fago, G., Marino, A., 2004. Analytical methodologies for determining the occurrence of endocrine disrupting chemicals in sewage treatment plants and natural waters. Anal. Chim. Acta 501 (1), 79–88 (Jan). Le Guevel, R., Pakdel, F., 2001. Assessment of oestrogenic potency of chemicals used as growth promoter by in-vitro methods. Hum. Reprod. 16 (5), 1030–1036 (May). Leffers, H., Naesby, M., Vendelbo, B., Skakkebaek, N.E., Jørgensen, M., 2001. Oestrogenic potencies of zeranol, oestradiol, diethylstilboestrol, bisphenol-A and genistein: implications for exposure assessment of potential endocrine disrupters. Hum. Reprod. 16 (5), 1037–1045 (May 1). Lehtinen, K.J., Mattsson, K., Tana, J., Engström, C., Lerche, O., Hemming, J., 1999. Effects of wood-related sterols on the reproduction, egg survival, and offspring of brown trout (Salmo trutta lacustris L.). Ecotoxicol. Environ. Saf. 42 (1), 40–49 (Jan). Leusch, F.D.L., MacLatchy, D.L., 2003. In vivo implants of β-sitosterol cause reductions of reactive cholesterol pools in mitochondria isolated from gonads of male goldfish (Carassius auratus). Gen. Comp. Endocrinol. 134 (3), 255–263 (Dec). Leusch, F.D.L., De Jager, C., Levi, Y., Lim, R., Puijker, L., Sacher, F., et al., 2010. Comparison of five in vitro bioassays to measure estrogenic activity in environmental waters. Environ. Sci. Technol. 3853–3860. Liggins, J., Bluck, L.J., Runswick, S., Atkinson, C., Coward, W.A., Bingham, S.A., 2000. Daidzein and genistein contents of vegetables. Br. J. Nutr. 84 (5), 717–725 (Nov). Liu, Z.H., Kanjo, Y., Mizutani, S., 2010. A review of phytoestrogens: their occurrence and fate in the environment. Water Res. 44 (2), 567–577. Lundgren, M.S., Novak, P.J., 2009. Quantification of phytoestrogens in industrial waste streams. Environ. Toxicol. Chem. 28 (11), 2318. MacLatchy, D., Peters, L., Nickle, J., van der Kraak, G., 1997. Exposure to beta-sitosterol alters the endocrine status of goldfish differently than 17beta-estradiol. Environ. Toxicol. Chem. 16 (9), 1895–1904 (Sep). Maragos, C.M., 2012. Zearalenone occurrence in surface waters in central Illinois, USA. Food Addit. Contam. Part B Surveill. 5 (1), 55–64 (Jan).

43

Massart, F., Saggese, G., 2010. Oestrogenic mycotoxin exposures and precocious pubertal development. Int. J. Androl. 369–376. Matsumura, A., Ghosh, A., Pope, G.S., Darbre, P.D., 2005. Comparative study of oestrogenic properties of eight phytoestrogens in MCF7 human breast cancer cells. J. Steroid Biochem. Mol. Biol. 94, 431–443. Matthiessen, P., Arnold, D., Johnson, A.C., Pepper, T.J., 2006. Contamination of headwater streams in the United Kingdom by oestrogenic hormones from livestock farms. 367, 616–630. Mattsson, K., Tana, J., Engström, C., Hemming, J., Lehtinen, K.J., 2001. Effects of woodrelated sterols on the offspring of the viviparous blenny, Zoarces viviparus L. Ecotoxicol. Environ. Saf. 49 (2), 122–130 (Jun). Metzler, M., Pfeiffer, E., Hildebrand, A.A., 2010. Zearalenone and its metabolites as endocrine disrupting chemicals. World Mycotoxin J. 3 (4), 385–401 (Nov). Minervini, F., Giannoccaro, A., Cavallini, A., Visconti, A., 2005. Investigations on cellular proliferation induced by zearalenone and its derivatives in relation to the estrogenic parameters. Toxicol. Lett. 159, 272–283. Molina-Molina, J.-M., Real, M., Jimenez-Diaz, I., Belhassen, H., Hedhili, A., Torné, P., et al., 2014. Assessment of estrogenic and anti-androgenic activities of the mycotoxin zearalenone and its metabolites using in vitro receptor-specific bioassays. Food Chem. Toxicol. 74C, 233–239 (Oct 16). Munkittrick, K.R., McMaster, M.E., McCarthy, L.H., Servos, M.R., Van Der Kraak, G.J., 1998. An overview of recent studies on the potential of pulp-mill effluents to alter reproductive parameters in fish. J. Toxicol. Environ. Health B Crit. Rev. 1 (4), 347–371. Murkies, A.L., Wilcox, G., Davis, S.R., 1998. Clinical review 92: phytoestrogens. J. Clin. Endocrinol. Metab. 83 (2), 297–303 (Mar). Nadzialek, S., Vanparys, C., Van der Heiden, E., Michaux, C., Brose, F., Scippo, M.-L., et al., 2010. Understanding the gap between the estrogenicity of an effluent and its real impact into the wild. Sci Total Environ. 408(4). Elsevier B.V., pp. 812–821 (Jan 15). Newill, H., Loske, R., Wagner, J., Johannes, C., Lorenz, R.L., Lehmann, L., 2007. Oxidation products of stigmasterol interfere with the action of the female sex hormone 17beta-estradiol in cultured human breast and endometrium cell lines. Mol. Nutr. Food Res. 51 (7), 888–898 (Jul). Panter, G.H., Hutchinson, T.H., Länge, R., Lye, C.M., Sumpter, J.P., Zerulla, M., et al., 2002. Utility of a juvenile fathead minnow screening assay for detecting (anti-)estrogenic substances. Environ. Toxicol. Chem. 21 (2), 319–326 (Feb). Pawlowski, S., Ternes, T., Bonerz, M., Kluczka, T., van der Burg, B., 2003. Combined in situ and in vitro assessment of the estrogenic activity of sewage and surface water samples. Toxicol. Sci. 57–65. Pelissero, C., Sumpter, J.P., 1992. Steroids and “steroid-like” substances in fish diets. Aquaculture 107 (4), 283–301 (Nov). Rearick, D.C., Fleischhacker, N.T., Kelly, M.M., Arnold, W.A., Novak, P.J., Schoenfuss, H.L., 2014. Phytoestrogens in the environment, I: occurrence and exposure effects on fathead minnows. Environ. Toxicol. Chem. 33 (3), 553–559 (Mar). Ribeiro, C., Pardal, M.A.Â., Martinho, F., Margalho, R., Tiritan, M.E., Rocha, E., et al., 2009a. Distribution of endocrine disruptors in the Mondego River estuary, Portugal. Environ. Monit. Assess. 149 (1–4), 183–193 (Feb). Ribeiro, C., Tiritan, M.E., Rocha, E., Rocha, M.J., 2009b. Seasonal and spatial distribution of several endocrine-disrupting compounds in the Douro River Estuary, Portugal. Arch. Environ. Contam. Toxicol. 56 (1), 1–11 (Jan). Rocha, M.J., Cruzeiro, C., Rocha, E., 2013. Development and validation of a GC–MS method for the evaluation of 17 endocrine disruptor compounds, including phytoestrogens and sitosterol, in coastal waters — their spatial and seasonal levels in Porto costal region (Portugal). J. Water Health 11 (2), 281–296 (Jun). Rochester, J.R., Millam, J.R., 2009. Phytoestrogens and avian reproduction: exploring the evolution and function of phytoestrogens and possible role of plant compounds in the breeding ecology of wild birds. Comp. Biochem. Physiol. A Mol. Integr. Physiol. 154 (3), 279–288 (Nov). Schenzel, J., Hungerbühler, K., Bucheli, T.D., 2012. Mycotoxins in the environment: II. Occurrence and origin in Swiss river waters. Environ Sci Technol. 46(24). American Chemical Society, pp. 13076–13084 (Dec 18). Schiller, V., Wichmann, A., Kriehuber, R., Muth-Kohne, E., Giesy, J.P., Hecker, M., et al., 2013. Studying the effects of genistein on gene expression of fish embryos as an alternative testing approach for endocrine disruption. Comp. Biochem. Physiol. C: Toxicol. Pharmacol. 41–53. Schmitt, S., Reifferscheid, G., Claus, E., Schlüsener, M., Buchinger, S., 2012. Effect directed analysis and mixture effects of estrogenic compounds in a sediment of the River Elbe. Environ. Sci. Pollut. Res. Int. 19 (8), 3350–3361 (Sep). Scholz, B., Liebezeit, G., 2012. Screening for biological activities and toxicological effects of 63 phytoplankton species isolated from freshwater, marine and brackish water habitats. Harmful Algae 20. Elsevier B.V., pp. 58–70 (Dec). Scholz, S., Kordes, C., Hamann, J., Gutzeit, H.O., 2004. Induction of vitellogenin in vivo and in vitro in the model teleost medaka (Oryzias latipes): comparison of gene expression and protein levels. Mar. Environ. Res. 57 (3), 235–244 (Apr). Schwartz, P., Thorpe, K.L., Bucheli, T.D., Wettstein, F.E., Burkhardt-Holm, P., 2010. Shortterm exposure to the environmentally relevant estrogenic mycotoxin zearalenone impairs reproduction in fish. Sci Total Environ. 409(2). Elsevier B.V., pp. 326–333 (Dec 15). Schwartz, P., Bucheli, T.D., Wettstein, F.E., Burkhardt-Holm, P., 2013. Life-cycle exposure to the estrogenic mycotoxin zearalenone affects zebrafish (Danio rerio) development and reproduction. Environ. Toxicol. 28 (5), 276–289 (May). Shappell, N.W., Mostrom, M.S., Lenneman, E.M., 2012. E-Screen evaluation of sugar beet feedstuffs in a case of reduced embryo transfer efficiencies in cattle: the role of phytoestrogens and zearalenone. In Vitro Cell. Dev. Biol. Anim. 48, 216–228. Smeds, A.I., Willför, S.M., Pietarinen, S.P., Peltonen-Sainio, P., Reunanen, M.H.T., 2007. Occurrence of “mammalian” lignans in plant and water sources. Planta 226 (3), 639–646.

44


Stackelberg, P.E., Furlong, E.T., Meyer, M.T., Zaugg, S.D., Henderson, A.K., Reissman, D.B., 2004. Persistence of pharmaceutical compounds and other organic wastewater contaminants in a conventional drinking-water-treatment plant. Sci. Total Environ. 329 (1–3), 99–113 (Aug). Stackelberg, P.E., Gibs, J., Furlong, E.T., Meyer, M.T., Zaugg, S.D., Lippincott, R.L., 2007. Efficiency of conventional drinking-water-treatment processes in removal of pharmaceuticals and other organic compounds. Sci. Total Environ. 377 (2–3), 255–272 (May). Stanford, B.D., Snyder, S.A., Trenholm, R.A., Holady, J.C., Vanderford, B.J., 2010. Estrogenic activity of US drinking waters: a relative exposure comparison. J. Am. Water Work Assoc. 102 (12), 55–65. Stevenson, L.M., Brown, A.C., Montgomery, T.M., Clotfelter, E.D., 2011. Reproductive consequences of exposure to waterborne phytoestrogens in male fighting fish Betta splendens. Arch. Environ. Contam. Toxicol. 60 (3), 501–510 (Apr). Sumpter, J.P., Johnson, A.C., 2008. 10th anniversary perspective: reflections on endocrine disruption in the aquatic environment: from known knowns to unknown unknowns (and many things in between). J. Environ. Monit. R. Soc. Chem. 1476–1485. Sun, Q., Deng, S., Huang, J., Shen, G., Yu, G., 2008. Contributors to estrogenic activity in wastewater from a large wastewater treatment plant in Beijing, China. Environ. Toxicol. Pharmacol. 25 (1), 20–26 (Jan).

Tremblay, L., Van der Kraak, G., 1999. Comparison between the effects of the phytosterol beta-sitosterol and pulp and paper mill effluents on sexually immature rainbow trout. Environ. Toxicol. Chem. 329–336. Wang, C., Wang, L., Zhao, Q., Chen, J., Zheng, L., Zheng, M., et al., 2013. Cloud point extraction coupled with HPLC-DAD for the determination of genistein and daidzein in river water. Anal. Methods 5 (15), 3688. Waśkiewicz, A., Gromadzka, K., Bocianowski, J., Pluta, P., Goliński, P., 2012. Zearalenone contamination of the aquatic environment as a result of its presence in crops. Arh. Hig. Rada Toksikol. 63 (4), 429–435 (Dec). Welshons, W.V., Rottinghaus, G.E., Nonneman, D.J., Dolan-Timpe, M., Ross, P.F., 1990. A sensitive bioassay for detection of dietary estrogens in animal feeds. J. Vet. Diagn. Invest. 2 (4), 268–273 (Oct). Whitten, P.L., Patisaul, H.B., 2001. Cross-species and interassay comparisons of phytoestrogen action. Environ. Health Perspect. 109, 5–20 (Mar). WHO, UNEP, 2013. State of the Science of Endocrine Disrupting Chemicals 2012pp. 1–296.

Comprehensive assessment of hormones, phytoestrogens, and estrogenic activity in an anaerobic swine waste lagoon.

Phytoestrogens and Mycoestrogens Induce Signature Structure Dynamics Changes on Estrogen Receptor α.

The estrogenic activity of certain phytoestrogens in the Siberian sturgeon Acipenser baeri.

Estrogenic activity and identification of potential xenoestrogens in a coking wastewater treatment plant.

Anti-estrogenic activity formation potential assessment and precursor analysis in reclaimed water during chlorination.

Arctic contaminants: sources, occurrence and pathways.

The potential health effects of dietary phytoestrogens.

anti-estrogenic activity?

Estrogenic and anti-estrogenic activity of off-the-shelf hair and skin care products.

Phytoestrogens in the environment, I: occurrence and exposure effects on fathead minnows.

Novel Promising Estrogenic Receptor Modulators: Cytotoxic and Estrogenic Activity of Benzanilides and Dithiobenzanilides.

The estrogenic and androgenic potential of pyrethroids in vitro. Review.

Potential contribution of dietary sources to urinary cadmium and beta 2-microglobulin excretion of occupationally exposed workers.

Global occurrence and emission of rotaviruses to surface waters.

Phytoestrogens and their effects.

Effects of beta-particles attenuation filter of reference sources and contribution in the calibration of surface contamination monitors.

MyD88 contribution to ocular surface homeostasis.

Occurrence and activity of anammox bacteria in surface sediments of the southern North Sea.

Estrogenic activity of o,p'-DDT metabolites and related compounds.

Aluminum and its potential contribution to Alzheimer's disease (AD).

Social neuroscience and its potential contribution to psychiatry.

Alternative RNA splicing: contribution to pain and potential therapeutic strategy.

Contribution by symbiotically luminous fishes to the occurrence and bioluminescence of luminous bacteria in seawater.

Confirmed and Potential Sources of Legionella Reviewed.