Effects of Estrogenic Compounds on Fish Populations in Fresh Water Aquatic Systems

Home , Atrazine, Nonylphenol

Tyler Saechao

California State University, Sacramento

Table of Contents

Abstract……………………………………………………………………………………………3

Introduction………………………………………………………………………………………..4

Methods……………………………………………………………………………………………5

Estrogenic Endocrine Disruption …………………………………………………………………5

Estrogen Source: Fate and Transport……………………………………………………………...6

Domestic…………………………………………………………………………………..6

Industrial…………………………………………………………………………………..7

Agriculture and Livestock…………………………………………………………………8

Fish Responses…………………………………………………………………………………….9

Bioaccumulation…………………………………………………………………………..9

Mechanism of Action…………………………………………………………………….10

Dose-Response…………………………………………………………………………………...11

Developmental Toxicity………………………………………………………………….12

Reproductive Toxicity…………………………………………………………………...13

Potential Population Decline……………………………………………………………..15

Identifying and Addressing Sources……………………………………………………………..16

Advanced Wastewater Treatment Plants………………………………………………...16

Discussion………………………………………………………………………………………..17

Figures and Tables……………………………………………………………………………19-24

References……………………………………………………………………………………25-28

2 Abstract

Many estrogenic compounds found in the environment are endocrine disruptors. This study looks at the effect of estrogenic endocrine disruptors on fish populations in aquatic environments. Sources of estrogenic compounds include households, industries, and farming operations. Negative impacts are development and reproductive, as teratogenicity and changes in morphology have been reported. Implementing advance procedures into waste water treatment plants have been shown to be effective in removing estrogenic compounds. Reducing the input of these compounds in surface water is crucial to maintaining health fish populations.

3 Introduction

Estrogenic compounds have become common in aquatic environments and known to negatively affect fish populations (Kidd et al., 2007; Moreman et al., 2017). Estrogenic compounds include estrogens and estrogen mimickers such as pharmaceuticals, industrial compounds, and chemicals used for farming (Markey et al. 2001; Fenet et al., 2003; Canesi &

Fabbri, 2015). There are several specific estrogenic compounds that result in adverse effects on fish.

Estrogenic compounds of particular significance include estrone (E1), estradiol (E2), estriol (E3), ethinyl-estradiol (EE2), bisphenol-A (BPA), nonylphenol, and atrazine (Markey et al., 2001; Goeppert et al., 2015). E1, E2, E3, and EE2 are pharmaceutical forms of estrogen

(Ebele et al. 2017). BPA and nonylphenol are common components in industry, while atrazine is a widely used pesticide (Markey et al. 2001; Tillitt et al., 2010). As these compounds are present in different operations, they are released into surface water, impacting fish (Amester, 2016;

Archer et al., 2017; Moore et al., 2011). These compounds have been found in freshwater systems and cause negative responses through endocrine disruption.

When impactful estrogenic compounds are in water, they bioaccumulate in fish.

Estrogens are hydrophobic lipid hormones and thus bioaccumulate in fish tissues (Christiansen et al. 2002). As fish are exposed to environmental estrogens their biology is altered (Schwindt et al., 2014; Wang et al., 2018). This occurs due to estrogen being an endocrine disruptor (Kidd et al., 2007 & Ward & Schoenfuss, 2017). Endocrine disruptors are exogenous substances that interfere with the endocrine system, the hormone system within living organisms (NIEHS,

2010). These chemicals disrupt the normal hormone secretion and adversely affect development and reproduction in wildlife. Similar to hormonal therapy, the exposure over time has been

4 shown to cause intersex in male fish and decrease the gender ratio in fish populations (Kidd et al., 2007; Harris et al., 2011; Ward & Schoenfuss, 2017). This has the potential to cause large declines of fish populations, as breeding will be reduced (Kidd et al., 2007; Ward & Schoenfuss,

2017). There are many studies that look at the effects of estrogens on fish, but not many that try to assess plans to address the source.

The primary goal of this study is to evaluate the effects of estrogens on aquatic environments. This study will look at synthetic estrogens, the different pathways in which these chemicals enter water, and their effects on fish. It will also examine current experiments that address sources, in an effort to find ways to minimize environmental impacts.

Methods

In this study, peer reviewed articles will be evaluated to address the negative impacts of estrogenic compounds on fish. This study will also include successfully implemented technology to reduce the amounts of environmental estrogens and endocrine disruptors. This will be done to address the issue of estrogenic compound contamination and reduce the input of these chemicals to surface water systems.

Estrogenic Endocrine Disruption

Estrogenic compounds are classified as endocrine disruptors because they affect the endocrine system in several ways (Figure 1). Endocrine disruptors are defined by the National

Institute of Environmental Health Sciences (NIEHS, 2010) as “natural occurring compounds or man-made substances that mimic or interfere with the function of hormones.” Disruptors can affect homeostasis by inhibiting estrogen production or by mimicking natural estrogens at the

5 receptors. This causes an overabundance of estrogen and results in overstimulation of the receptors (NIEHS, 2010; Pinto et al., 2014). As overstimulation of estrogens, this can cause several adverse effects such as male fish to develop female specific organs. This transpires due to the compound’s ability to mimic natural hormones within the body. They can also decrease the amount of estrogens by blocking normal estrogen activities resulting in deprivation of receptors or by altering the function of estrogen receptors and changing the synthesis, transport, metabolism, and excretion of hormones. As these effects occur, homeostasis provided by the endocrine system is altered which leads to detrimental effects.

Estrogen Sources: Fate and Transport

Estrogenic compounds through enter water systems through several different routes, which include households, industry, agriculture, and livestock (Figure 2) (Amester, 2016; Archer et al., 2017; Moore et al., 2011). Waste water treatment plants (WWTPs) are used to reduce contaminants in water for eventual discharge into the environment; however, these facilities are not sufficient enough to prevent estrogen compounds from infiltrating water (Archer et al., 2017;

Moore et al., 2011). As these chemicals all enter surface waters through WWTPs, they are is discharged back into surface waters.

Domestic Sources

Estrogens that come from households include pharmaceuticals like birth control and steroid hormones. The active substance within birth control that is found in surface water is ethinyl-estradiol (EE2). Steroid hormones include estrone (E1), estradiol (E2), and estriol (E3)

(Goeppert et al., 2015). The pharmaceutical forms of estrogen typically enter surface waters

6 through improper disposal. As the public disposes these medications through waste drains, toilets, or in the trash, they end up in WWTPs and landfills. (Amster, 2016; Archer et al., 2017).

These compounds also enter WWTPs as they are metabolized and excreted. Once urine and feces are flushed down the toilet, they transport estrogens into WWTPs (Moore et al, 2011; Archer et al., 2017).

Worldwide discharge rates of E1, E2, and E3 has been estimated to be discharged around rates of 30,000 kg per year, while EE2 is around 700 kg per year (Pengchen et. al, 2018). In the

United States, sewage effluent and surface water have reported levels of E2 and EE2. E2 concentrations in effluent and surface water has been found between 0.1 to 4.05 and 0.107 to

2.67, respectively (Christiansen et al. 2002). Concentrations of EE2 in effluent and surface water has been found between 0.1 to 2.42 and 0.053 to 0.52 ng/L, respectively (Christiansen et al.

2002).

Once discharged into surface water, estrogens have shown to biodegrade. E1 and E2 has a half life between 0.2 to 8.7 days and 0.1 to 10.9 days in conditions that are aerobic

(Christiansen et al. 2002). E1 and E2 biodegrade quick in comparison to EE2. The half-life for

EE2 in aerobic conditions is ten times greater than E1, taking 17 days (Christiansen et al. 2002).

This shows that EE2 is more persistent than other steroidal estrogens.

Industrial Pollution

Industries discharges endocrine disruptors both directly into the environment and indirectly to through WWTPs. There are several compounds that have the ability to mimic estrogens and disrupt the endocrine system, such as BPA and alkylphenols (Lindholst et al.,

2006; Markey et al., 2001). BPA is a very common substance in manufactured products, as it is

7 used as starting material in polycarbonate plastics (i.e. water bottles, trash linings, etc.), paper, and metals (US EPA, 2010). Nonylphenols are used for manufacturing detergents, polymers, antioxidants, etc. These compounds enter surface water through WWTPs (Ying et al. 2002). As estrogen mimickers, they also contribute to endocrine disruption in aquatic wildlife. BPA waste produced at manufacturing companies enter WWTPs (Markey et al., 2001; Canesi & Fabbri,

2015).

In the United States, concentrations of BPA in effluent and surface water has been reported between 0.02 to 0.055 ug/L and 1 to 8 ug/L, respectively (Christiansen et al. 2002).

Concentrations of nonylphenols in effluent and surface water has been reported between 1 to 33 ug/L and 0.11 to 0.64 ug/L, respectively (Christiansen et al. 2002). The half-life of BPA (2.5 to 4 days) and nonylphenols (7 to 28 days) in surface water was found (Christiansen et al. 2002).

Agriculture and Livestock

Products used for farming operations, such as pesticides and fertilizers, have been shown to contain estrogenic compounds and which can enter surface water (Tillit et al. 2010). For example, atrazine is a widely used herbicide in the U.S. is has been determined to be an estrogen mimicker (Tillit et al. 2010). As this herbicide is applied to land to control weeds, runoff occurs, and the compound enters surface and ground waters.

Livestock operations are a source of water contamination. Hormones are widely used in livestock productions. These growth hormones are synthetic estrogens that are used to enhance the growth of cattle, swine, chickens, and ducks (IFATP, 2009). This contributes to the presence of estrogenic compounds in animal waste. A study that analyzed the composition of livestock manure found concentrations of both BPA and EE2 (Pengchen et. al, 2018). The manure is

8 applied to crops to promotes crop growth. Eventually leaving the fields through runoff, the compounds into natural waters (Lucas & Jones, 2006). As evidence through manure, livestock productions contribute to estrogen pollution. As these animals produce waste, runoff occurs which allows the compounds to enter water.

The United States has an estimated discharge rates of 83,000 kg/year from livestock

(Pengchen et. al, 2018). These higher rates may be due concentrated animal feeding operations

(CAFOs) and the fact that these compounds enter water without treatment. CAFOs are significant sources of estrogenic compounds, but since these compounds enter water through runoff, quantifying the rates of pollution is difficult to determine (Pengchen et. al, 2018;

Christiansen et al. 2002). These compounds have been reported to enter the environment at low levels, but wildlife is considered highly susceptible to these discharges.

Fish Responses

Bioaccumulation

As common estrogenic compounds enter the water and fish are exposed, they bioaccumulate within the organism. Bioaccumulation is defined by the United States

Environmental Protection Agency (U.S. EPA, 2012) as chemicals “taken up by a plant or animal either directly from exposure to a contaminated medium (soil, sediment, water) or by eating food containing the chemical”. Fish are rich in fat, while allows highly lipophilic compounds like estrogens to readily bioaccumulate. While these compounds do biodegrade over time, most of them are very persistent within water (Christiansen et al. 2002). The persistence of these compounds displays a great potential for bioaccumulation.

9 These compounds can enter fish bodies through aquatic respiration, dermally, and through the consumption of other organisms (biomagnification). In a study done along the Bahe

River located in China, surface water and fresh fish samples were collected (Wang et al. 2018).

This was done in an attempt to determine the bioconcentration factor (BCF), which is a ratio of the concentration of the compounds present within the fish in relation to surrounding environment (Table 1). Using the highest level of compounds detected within …, the BCF was determined for nonylphenol (0.16 g L), bisphenol-A (0.09 g L), estrone (0.25 g L), estradiol

(0.38 g L), ethinyl-estradiol (0.43 g L), and estriol (0.25 g L), as concentrations within fish muscle tissue over concentrations in surface water (Wang et al. 2018). This displays the compounds potential to concentrate within the tissues of a living organism. The concentrations of these chemicals can be found through the fish body, but with adverse effects occurring due to the interaction with the hypothalamus, pituitary, and gonadal axis.

Mechanism of Action

As xeno-estrogens are taken into the body of fish, they begin to affect the fish on a cellular level within their tissue and the hypothalamus-pituitary-gonadal axis. Estrogenic endocrine disruptors can to act at several levels (Figure 3) as an agonist or antagonist, which contribute to negative impacts on fish. Estrogen receptors are widespread and found within nuclei, cytoplasm, and other organelles. Estrogenic compounds have been shown to regularly bind to the intracellular estrogen receptors (ERα, ERβa or ERβb) within fish (Ropero et al.,

2006; Pinto et al., 2014). As these receptors are activated, they can then bind to DNA via estrogen response elements or cause indirect genomic effects through different transcription factors (Ropero et al., 2006; Pinto et al., 2014).

10 These compounds can also cellular affect fish, as they bind to plasma membrane estrogen receptors. Upon membrane activation, protein kinase cascade occurs causing transcription factors to activate resulting in alteration of gene expression (Ropero et al., 2006; Pinto et al., 2014). This transcription factor activation can occur through the effects of specific enzymes that occur via G- protein coupled receptors (GPER). This process takes place as adenylyl cyclase is initiated, inhibiting cAMP production and thus the stimuli PKA, resulting in rapid non-genomic effects

(Ropero et al., 2006; Pinto et al., 2014). The alteration to DNA results in adverse effects expressed in fish.

The hypothalamic–pituitary–gonadal axis is responsible for promoting healthy development, reproduction, and behavior. These compounds cause several processes to occur which include affecting the expression of CYP19A aromatase and the production of vitellogenin in male fish (Page et al., 2011). As homeostatic hormone levels are disrupted, the species can experience teratogenic effects and impotence, which results in low fecundity (Schwindt et al.,

2011).

Dose-Response

Fish are one of the most sensitive organisms to estrogenic endocrine disruptors. These compounds have been shown to adversely affect development and reproduction. The lowest observed adverse effect level (LOAEL) was determined for estrogenic compounds and was characterized by the lowest concentration at which VTG production occurs. The LOAEL of EE2 in experimental fish populations ranged from 0.1 to 5 ng/L (Christiansen et al. 2002; Young et al., 2002). Nonylphenol has a LOAEL of 1 ug/L for rainbow trout (Fent et al., 2000). BPA has a

LOAEL between 40 to 70 ug/L for rainbow trout (Lindholst et al., 2000). The lethal dose for

11 50% (LD50) of fish embryos for EE2 and BPA was determined to be 5 ug/L and 11.84 mg/L, respectively (Ortiz-Zarragoitia et al., 2009; Moreman et al., 2017). Without a healthy reproductive system, this can result in the collapse of fish populations. The effects of endocrine disruptors among individual fish have been reported as mild to severe.

Developmental Toxicity

Endocrine disruptors have been shown to cause adverse effects in development

(teratogenic effects) and reproduction in fish. Teratogenicity can be characterized as birth defects and greater potential for loss of embryo. Several studies have looked at these adverse effects on fish.

Estrogenic endocrine disruptors have the ability to cause delays in fish development and cause greater rates of mortality. Both EE2 and BPA have been shown to cause adverse effects on fish development. Embryos that were exposed to doses of EE2 at 5 µg/L over the course of 10 days post fertilization, experienced a 50% mortality and a delay in development (Ortiz-

Zarragoitia et al., 2009). An experiment was conducted on zebrafish (Danio rerio) embryos from

0 to 96 hours post fertilization (hpf), in which bisphenols were applied to controlled environments (Moreman et al., 2017). BPA was shown to cause delay in hatching and at a dose of 11.84 mg/L displayed 50% mortality of fish larvae (Figure 4). The BPA induced malformations observed in this experiment include cardiac edema (≥ 5.0 mg/L), craniofacial malformation (≥ 10.0 mg/L), cranial hemorrhage (≥ 12.5 mg/L, and yolk sac deformations (≥

10.0 mg/L) (Figure 5). Exposure has been shown to cause changes to the normal biological composition in fish.

12 Two studies looking at estrogenic endocrine disruptors on development noted structural changes to fish skeletons. Both were in vitro studies on fish species conducted during early life stages, albeit different species. One of the studies was conducted on the fathead minnow

(Pimephales promelas), during a period of 24 hpf to 26 days post-hatch (Figure 6). EE2 and BPA were applied at 0.1 to 100 µg/L and 1000 µg/L, respectively. While BPA showed no significant changes to vertebrae structure, EE2 showed 62% of the fish within the study to having skeletal malformations (Warner & Jenkins, 2007). This change in morphology could cause significant issues in wildlife populations, as reproductive successes has been linked to body length (Harris et al. 2011).

Another study showed significant skeletal malformations from xeno-estrogens on fish.

The experiment was conducted on estuary mummichogs (Fundulus heteroclitus), in which EE2 produced significant changes at 1000 ng/L doses, applied from 25 to 60 days post hatch

(Boudreau et al., 2004). The observed results showed both skeletal and soft tissue anomalies with an increased percentage of malformations in fish. EE2 has the ability to cause delays in fish development and cause greater rates of mortality. These malformations have the potential to reduce reproductive success (Harris et al., 2011).

Reproductive Toxicity

The most significant adverse effect of estrogens on fish is the feminization of their populations. The production of yolk protein specific to female fish during oocyte maturation, vitellogenin (VTG), is an indicator of exposure to environmental estrogens in male fish (Hansen et al.,1998; Schultz et al., 2000; Christiansen et al. 2002). The lowest observed adverse effect level (LOAEL) is indicated as the lowest dose in which vitellogenin production occurs but

13 differs depending on the fish species. The LOAEL of EE2 has been reported at 0.1 ng/L for the

Japanese medaka (Oryzias latipes) and 1-5 ng/L for rainbow trout (Oncorhynchus mykiss) and zebra fish (Christiansen et al. 2002; Young et al. 2002). EE2 has been shown to have the highest potency. The LOAEL of nonylphenols for rainbow trout is 1 ug/L during long term exposure

(Fent et al., 2000). The LOAEL of BPA for rainbow trout is from 40-70 ug/L (Lindholst et al.,

2000). This overall effect has the ability to affect gender ratio among fish and cause intersex males.

The feminization of fish is manifested through increased females in fish populations and intersex of male fish. Intersex in male fish is characterized as the production of both testis and ovary tissues. Due to this, gender ratios are also significantly changed. This effect reduces the fertility of male fish as well as the success of reproduction. During in vivo assays on rainbow trout and Japanese medaka, EE2 has been shown to cause intersex between dose of 10-100 ng/L

(Schultz et al., 2000; Metcalfe et al., 2001). At concentrations of 100 ng/L, EE2 caused all

Japanese medaka males to display testis-ova and 96% of female fish within population (Metcalfe et al., 2001). Nonylphenols cause intersex of Japanese medaka at 50 ug/L and 100 ug/L, affecting 50% and 86% of subjects, respectively (Metcalfe et al., 2001). The gender ratio within the control was 2 males to 1 female, but with the addition of nonylphenols at 100 ug/L, it declined (1M/2F) (Metcalfe et al., 2001). BPA, over a long period of exposure, has been shown to cause intersex at 10 ug/L (Metcalfe et al., 2001).

Within aquatic environments, the feminization of fish negatively impacts breeding. The capacity of the fish to reproduce varies, depending on the severity of feminization. A study that conducted two experiments, looking at the success of reproduction at different degrees of feminization, mild and severe (Harris et al, 2011). The severity was categorized through the

14 number of oocytes within the testes (Harris et al., 2011). In the mildly feminized fish (<2 oocytes), body length differences male and female fish resulted in a decline in reproductive success. In the severely feminized fish (≥ 4 oocytes), researchers found a negative correlation between reproduction and severe intersex (Harris et al., 2011). This was attributed to a reproduction decline of 76% among severely feminized individuals. These effects hindered the ability of successful reproduction at different exposure levels (Harris et al, 2011). Long term exposure to estrogens within their native habitat can cause significant damage to wild fish populations.

Potential Population Decline

As estrogenic endocrine disruptors enter natural waters, entire fish populations can be affected. The levels of these compounds in surface waters have been reported to be low, but still have the potential to cause decline in populations (Kidd et al. 2007). While there are no current reported instances of estrogenic compounds having affected wild populations, studies have been done to simulate possible outcomes.

An experiment was conducted on a whole lake (Experimental Lake Area 260) in Ontario,

Canada observing the effects of estrogens on fathead minnows (Pimephales promelas) over the course of 7 years (Figure 7). After 2 years of monitoring, EE2 was added at 5 ng L-1 to replicate relevant concentrations discharged by sewage treatment plants for 3 years, during open water season and the effects on fish were recorded (Kidd et al. 2007). Both male and female fish exhibited higher levels of VTG compared to the control during the 3 years (Figure 8). During spring 2002, the year after first EE2 amendments, males exhibited spermatogenesis delays, tissue damage, and tubule malformations and female fish exhibited delay of ovarian development

15 (Kidd et al. 2007). During this same year, reproductive success significantly declined, and population size decreased (Figure 9). After 2003, male fathead minnows began to display ova- testes and female fish showed an increase of atretic follicles within ovaries (Kidd et al. 2007).

Due to the imbalance of gender ratio and alteration of biology, the fish population had drastically declined. This shows estrogenic endocrine disruptors can cause detrimental effects to wildlife sustainability. With the consequences to wildlife being severe, reducing the contaminants is essential to health of aquatic species.

Identifying and Addressing Sources

One of the primary ways estrogenic endocrine disruptors enter surface waters is through

WWTPs. Wastewater includes domestic, industry, and agriculture waste. They are useful for negating potentially toxic substances in water, but the current standard of technology is not sufficient in filtering out estrogenic compounds. The process typically starts with a pretreatment followed by the three main stages of treatment: primary, secondary, and tertiary. The processes are physical, biological, and chemical, respectively. Effluent is this discharged back into surrounding waterbodies such as streams, rivers, and lakes. As this occurs, organisms such as fish are exposed. Studies have been conducted to look at the effectiveness of adopting new processes in the treatment to reduces estrogen pollution.

Advanced Wastewater treatment plants

Implementing additional technology to WWTPs is necessary in minimizing estrogenic compounds in the environment. Advanced procedures have been tested using chlorine dioxide

(ClO2), granular activated carbon (GAC), powdered activated carbon dust (PAC), and

16 nanofiltration (NF) to reduce compound levels and prevent feminization of fish. A study done in the Poland tested the effectiveness of GAC and PAC, coagulation, and NF (Bodzek & Dudziak

2006). PAC and GAC was found to be the most efficient in reducing the estrogens in effluents, followed by NF and coagulation (Bodzek & Dudziak 2006). The effectiveness of removal was tested against coagulates and looked at the percentage of removal of estrogenic compounds

(Figure 10). Sorption was determined to be the most efficient in removing compounds. This shows the benefits in adopting additional processes within current sewage treatment plants to remove these contaminants.

A study in Korea, tested the level of effectiveness between different types granular activated carbon (coal, coconut, and wood) on endocrine disruptors (Choi et al. 2005). The researchers observed that compounds with higher octanol-water partition coefficient (Kow) has greater sorption by GACs. As nonylphenols (5.76) has a higher Kow, it was better adsorbed than bisphenol-A (3.32) (Choi et al. 2005). The rate of removal was found to be greatest with coal- based carbon for both nonylphenols and BPA (Choi et al. 2005). The lowest rate of removal for nonylphenols and BPA was coconut and wood-based carbon, respectively. This is due to pore volume of the different forms of GAC, as higher adsorption is connected to larger pore volume

(Choi et al. 2005). This shows the usefulness of implementing GAC to reduce estrogen mimickers.

Discussion

Estrogenic compounds are harmful to fish. These compounds enter water through various sources including households, industries, agriculture, and livestock through WWTP and landfills

(Moore et al., 2011; Amester, 2016; Archer et al., 2017). Though the levels of compounds

17 entering surface water are low, they still bioaccumulate within fish species (Christiansen et al.,

2002; Wang et al. 2018). These compounds act at multiple estrogen receptors within fish, exhibiting several adverse effects. This includes cardiac edema, spinal malformation, craniofacial abnormalities, cranial hemorrhage, and yolk sac malformation (Mooreman et al.,

2017). Reproductive toxicity also occurs as male fish become intersex and the gender ratio of the population is altered. As this occurs populations of fish have the potential to experience declines.

The need for adopting new procedures and technology in current WWTPs is necessary for addressing these issues. WWTPs are the first line of defense for aquatic wildlife and the effectiveness of WWTPs is essential for the health over the fish populations and overall freshwater ecosystem. PAC and GAC have been shown to be the most efficient in removing estrogenic compounds from effluent (Bodzek & Dudziak 2006). While most of the studies on fish population declines have been experimental, preemptively implementing additional processes will prevent the collapse of fish populations (Kidd et al., 2007). As these compounds enter water through human activity (medication, industry, and farming), creating new policies and informing the public is essential in reducing these compounds.

18 Figures and Tables

(Figure 1. Processes by which endocrine disruptors can affect estrogen receptors (NIEHS, 2010))

Compound Tissue (ng g−1) Water (ng L-1) BCF

Nonylphenol 103.5 634.8 0.16304348

Bisphenol-A 146.9 1573.1 0.09338249

Estrone 14.2 55.9 0.25402504

Estradiol 9.3 23.9 0.38912134

Ethinyl-Estradiol 13.8 31.5 0.43809524

Estriol 1.3 5.2 0.25

(Table 1. Bioconcentration Factor (BCF) in the wild sharpbelly fish (Wang et al. 2018))

(Figure 2. Different pathways in which estrogenic compounds can enter water (Moore et al.

2011)

(Figure 3. Displays the different plays mechanisms of action for natural estrogens and endocrine disruptors can take place in fish (Pinto et al. 2014))

(Figure 4. A dose-response curve of BPA concentration and mortality of fish embryo/larvae (Mooreman et al., 2017))

(Figure 5. Image of the teratogenic effects observe A=normal, B=pigment lightening, C=cardiac edema, D=spinal malformation, E=craniofacial abnormalities, F=cranial hemorrhage, G=yolk sac malformation (Mooreman et al., 2017))

(Figure 6. Shows the exposure effects of EE2 on fathead minnows at different concentration over the course of 24 hpf to 26 days post hatch (Warner & Jenkins, 2007))

(Figure 7. Experimental site, analyzing the effects of environmentally relevant EE2 levels and their effects on fathead minnows (Kidd et al. 2007))

(Figure 8. Levels of VTG on female (top) and male (bottom) fathead minnows compared to reference site (Kidd et al. 2007))

(Figure 9. Frequency of fathead minnows using trap nets in Experimental Site 260 over Fork length (tip of snout to tail) A. Reference site B. EE2 amendments (Kidd et al. 2007))

(Figure 10. Percent removal of estrogens in water, using different methods A. Coagulation B. Sorption, (Bodzek & Dudziak 2006))

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