<td>Under development: Not open for comment. Do not cite</td>
<td>Under Development: Contributions and Comments Welcome</td>
<td></td>
<td>1.29</td>
<td>Under Development</td>
</tr>
</tbody>
</table>
</div>
</div>
<!-- Abstract Section, text as generated by author -->
<div id="abstract">
<h2>Abstract</h2>
<p>This AOP describes the linkages between agonism of the estrogen receptor (ER) and population relevant impacts on reproductive function in a range of oviparous vertebrates including amphibia, birds and fish. The information in this AOP for ER agonism does not apply to mammalian species and also not to invertebrates.</p>
<p> </p>
<p><br />
Amphibians are sensitive to ER agonists during the transformation from larval tadpole to juvenile frog as these include critical periods of metamorphic development and sex differentiation that may be particularly sensitive to endocrine disruption. Larvae exposed to ER agonists during mid-metamorphosis show developmental effects, a subsequent strong female-biased sex ratio which suggests that transient early life-stage exposure to ER agonists can produce effects on the reproductive organs that persist into the beginning of adult life-stages. Birds are also known to be vulnerable to ER agonists causing disruption of estrogen-regulated functions such as sexual differentiation and sexual behaviour. Model species such as the Japanese quail have been widely used as a model for studying various long-term effects after embryonic exposure to ER agonists. In terms of teleost fish, exposure to ER agonists leads to a suite of adverse outcomes depending upon whether exposures occur during or beyond the larval, juvenile and adult life-stages. For example, aquatic exposure to potent ER agonists during the larval and juvenile life-stages may leads to gonadal and renal pathology and skewed-sex ratios in adult fish (potentially 100% females). Larval, juvenile and adult male fish exposed to the same ER agonists display abnormal plasma or whole body levels of vitellogenin (VTG). Cumulative fecundity in adult populations is also adversely affected by ER agonists and this is an important endpoint in the OECD Test Guideline 229 Fish Short Term Reproduction Assay. In summary, this AOP has utility in supporting the application of test methods for detecting ER agonists, or in silico predictions of the ability of chemicals to act as ER agonists and cause impaired sexual development and reproductive dysfunction.</p>
<h2>Abstract</h2>
<hr>
<p>This AOP describes the linkages between agonism of the estrogen receptor (ER) and population relevant impacts on reproductive function in a range of oviparous vertebrates including amphibia, birds and fish. The information in this AOP for ER agonism does not apply to mammalian species and also not to invertebrates.
</p><p><br />
</p><p><br />
Amphibians are sensitive to ER agonists during the transformation from larval tadpole to juvenile frog as these include critical periods of metamorphic development and sex differentiation that may be particularly sensitive to endocrine disruption. Larvae exposed to ER agonists during mid-metamorphosis show developmental effects, a subsequent strong female-biased sex ratio which suggests that transient early life-stage exposure to ER agonists can produce effects on the reproductive organs that persist into the beginning of adult life-stages. Birds are also known to be vulnerable to ER agonists causing disruption of estrogen-regulated functions such as sexual differentiation and sexual behaviour. Model species such as the Japanese quail have been widely used as a model for studying various long-term effects after embryonic exposure to ER agonists. In terms of teleost fish, exposure to ER agonists leads to a suite of adverse outcomes depending upon whether exposures occur during or beyond the larval, juvenile and adult life-stages. For example, aquatic exposure to potent ER agonists during the larval and juvenile life-stages may leads to gonadal and renal pathology and skewed-sex ratios in adult fish (potentially 100% females). Larval, juvenile and adult male fish exposed to the same ER agonists display abnormal plasma or whole body levels of vitellogenin (VTG). Cumulative fecundity in adult populations is also adversely affected by ER agonists and this is an important endpoint in the OECD Test Guideline 229 Fish Short Term Reproduction Assay. In summary, this AOP has utility in supporting the application of test methods for detecting ER agonists, or in silico predictions of the ability of chemicals to act as ER agonists and cause impaired sexual development and reproductive dysfunction.
</p>
<br>
</div>
<!-- Background Section, text as generated by author -->
<div id="background">
<br>
</div>
<!-- AOP summary, includes summary of each of the events associated with this aop -->
<!-- Overall assessment section, *** what is included here? *** -->
<div id="overall_assessment">
<h2>Overall Assessment of the AOP</h2>
<p>In terms of the criteria associated with Key Events in this AOP, the following observations have been made as shown in parentheses []:
</p><p>1. concordance of dose-response relationships?; [There is strong dose-response relationship concordance over a wide range of experimental studies using ER agonists in well-defined animals models, including amphibians, birds and fish];
</p><p>2. temporal concordance among the key events and adverse effect?; [There is strong temporal concordance from partial and full life-cycle studies using ER agonists in well-defined animals models];
</p><p>3. strength, consistency, and specificity of association of adverse effect and initiating event?; [In fish, there is a strong and consistent association between ER agonist exposure, disruption of sexual development and reproductive dysfunction. The same is true for amphibians and birds although the published studies are less numerous.];
</p><p>4. biological plausibility, coherence, and consistency of the experimental evidence?; [For the oviparous species frequently studied to date, there is a high level of biological plausibility, coherence, and consistency across the published experimental evidence];
</p><p>5. alternative mechanisms that logically present themselves and the extent to which they may distract from the postulated AOP?; [Other mechanisms of relevance to estrogen-mediated sexual development include the disruption of the steroidogenic pathways (eg see the AOP for aromatase inhibition in fish) and this alterative AOP should be considered alongside ER agonism in the context of elevated plasma VTG levels, disrupted sexual development of reproductive dysfunction. The possibility of other AOPs arisign should be kept in mind through critical analysis of the updated pree-reviewed literature];
</p><p>6. uncertainties, inconsistencies and data gaps?; [An important aspect of uncertainty is quantifying the degree to which disrupted sexual development leads to a population-relevant impact via reproductive dysfunction. Experimental and validated population modelling is a key need to address this data gap and uncertainty. In the author's view, there are no major scientific inconsistencies with regard to the ER agonism AOP and associated Key Events].
<p>In terms of the criteria associated with Key Events in this AOP, the following observations have been made as shown in parentheses []:</p>
<p>1. concordance of dose-response relationships?; [There is strong dose-response relationship concordance over a wide range of experimental studies using ER agonists in well-defined animals models, including amphibians, birds and fish];</p>
<p>2. temporal concordance among the key events and adverse effect?; [There is strong temporal concordance from partial and full life-cycle studies using ER agonists in well-defined animals models];</p>
<p>3. strength, consistency, and specificity of association of adverse effect and initiating event?; [In fish, there is a strong and consistent association between ER agonist exposure, disruption of sexual development and reproductive dysfunction. The same is true for amphibians and birds although the published studies are less numerous.];</p>
<p>4. biological plausibility, coherence, and consistency of the experimental evidence?; [For the oviparous species frequently studied to date, there is a high level of biological plausibility, coherence, and consistency across the published experimental evidence];</p>
<p>5. alternative mechanisms that logically present themselves and the extent to which they may distract from the postulated AOP?; [Other mechanisms of relevance to estrogen-mediated sexual development include the disruption of the steroidogenic pathways (eg see the AOP for aromatase inhibition in fish) and this alterative AOP should be considered alongside ER agonism in the context of elevated plasma VTG levels, disrupted sexual development of reproductive dysfunction. The possibility of other AOPs arisign should be kept in mind through critical analysis of the updated pree-reviewed literature];</p>
<p>6. uncertainties, inconsistencies and data gaps?; [An important aspect of uncertainty is quantifying the degree to which disrupted sexual development leads to a population-relevant impact via reproductive dysfunction. Experimental and validated population modelling is a key need to address this data gap and uncertainty. In the author's view, there are no major scientific inconsistencies with regard to the ER agonism AOP and associated Key Events].</p>
<a href="#Sex_Applicability"> Sex Applicability</a><br />
In terms of the taxonomic domains of applicability, exposure to ER agonists is capable of disrupting sexual development and causing reproductive dysfunction in oviparous species suchas amphibians, birds and fish (see examples of peer-revised literature cited below).
In terms of the taxonomic domains of applicability, exposure to ER agonists is capable of disrupting sexual development and causing reproductive dysfunction in oviparous species suchas amphibians, birds and fish (see examples of peer-revised literature cited below).</p>
</div>
<!-- potential consierations, text as entered by author -->
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<div id="references">
<h2>References</h2>
<hr>
<p><br />
Dang, Z., Traas, T., Vermeire, T. (2011) Evaluation of the fish short term reproduction assay for detecting endocrine disrupters. Chemosphere 85: 1592-1603
</p><p>Halldin, K., Axelsson, J., Brunström, B., (2005) Effects of endocrine modulators on sexual differentiation and reproductive function in male Japanese quail. Brain Research Bulletin 65: 211-218
</p><p>Hogan, N.S., Duarte, P., Wade, M.G., Lean, D.R.S., Trudeau, V.L. (2008) Estrogenic exposure affects metamorphosis and alters sex ratios in the northern leopard frog (Rana pipiens): Identifying critically vulnerable periods of development. General and Comparative Endocrinology 156: 515-523
</p><p>Hutchinson T.H. (2002) Impacts of endocrine disrupters on fish development: opportunities for adapting OECD Test Guideline 210. Environmental Sciences 9: 439-450
</p><p>Länge R., Hutchinson T.H., Croudace C.P., Siegmund F., Schweinfurth H., Hampe P., Panter G.H., Sumpter J.P. (2001) Effects of the synthetic oestrogen 17-ethinylestradiol over the life-cycle of the fathead minnow. Environmental Toxicology and Chemistry 20: 1216–1227
</p><p>Leino, R.L., Jensen,K.M., Ankley, G.T. (2005) Gonadal histology and characteristic histopathology associated with endocrine disruption in the adult fathead minnow (Pimephales promelas). Environmental Toxicology and Pharmacology 19: 85-98
</p><p>Ottinger, M.N., Carro, T., Bohannon, M., Baltos,L., Marcell, A.M., McKernan, M., Dean, K.M., Lavoie, E., Abdelnabi, M. (2013) Assessing effects of environmental chemicals on neuroendocrine systems: Potential mechanisms and functional outcomes. General and Comparative Endocrinology 190: 194-202
</p>
<br>
<p><br />
Dang, Z., Traas, T., Vermeire, T. (2011) Evaluation of the fish short term reproduction assay for detecting endocrine disrupters. Chemosphere 85: 1592-1603</p>
<p>Halldin, K., Axelsson, J., Brunström, B., (2005) Effects of endocrine modulators on sexual differentiation and reproductive function in male Japanese quail. Brain Research Bulletin 65: 211-218</p>
<p>Hogan, N.S., Duarte, P., Wade, M.G., Lean, D.R.S., Trudeau, V.L. (2008) Estrogenic exposure affects metamorphosis and alters sex ratios in the northern leopard frog (Rana pipiens): Identifying critically vulnerable periods of development. General and Comparative Endocrinology 156: 515-523</p>
<p>Hutchinson T.H. (2002) Impacts of endocrine disrupters on fish development: opportunities for adapting OECD Test Guideline 210. Environmental Sciences 9: 439-450</p>
<p>Länge R., Hutchinson T.H., Croudace C.P., Siegmund F., Schweinfurth H., Hampe P., Panter G.H., Sumpter J.P. (2001) Effects of the synthetic oestrogen 17-ethinylestradiol over the life-cycle of the fathead minnow. Environmental Toxicology and Chemistry 20: 1216–1227</p>
<p>Leino, R.L., Jensen,K.M., Ankley, G.T. (2005) Gonadal histology and characteristic histopathology associated with endocrine disruption in the adult fathead minnow (Pimephales promelas). Environmental Toxicology and Pharmacology 19: 85-98</p>
<p>Ottinger, M.N., Carro, T., Bohannon, M., Baltos,L., Marcell, A.M., McKernan, M., Dean, K.M., Lavoie, E., Abdelnabi, M. (2013) Assessing effects of environmental chemicals on neuroendocrine systems: Potential mechanisms and functional outcomes. General and Comparative Endocrinology 190: 194-202</p>
<td><a href="/aops/29">Aop:29 - Estrogen receptor agonism leading to reproductive dysfunction</a></td>
<td>MolecularInitiatingEvent</td>
</tr>
<tr>
<td><a href="/aops/52">Aop:52 - ER agonism leading to skewed sex ratios due to altered sexual differentiation in males</a></td>
<td>MolecularInitiatingEvent</td>
</tr>
<tr>
<td><a href="/aops/53">Aop:53 - ER agonism leading to reduced survival due to renal failure</a></td>
<td>MolecularInitiatingEvent</td>
</tr>
<tr>
<td><a href="/aops/536">Aop:536 - Estrogen receptor agonism leading to reduced survival and population growth due to renal failure</a></td>
<td>MolecularInitiatingEvent</td>
</tr>
<tr>
<td><a href="/aops/537">Aop:537 - Estrogen receptor agonism leads to reduced fecundity via increased vitellogenin in the liver</a></td>
<td>MolecularInitiatingEvent</td>
</tr>
<tr>
<td><a href="/aops/112">Aop:112 - Increased dopaminergic activity leading to endometrial adenocarcinomas (in Wistar rat)</a></td>
<td>KeyEvent</td>
</tr>
<tr>
<td><a href="/aops/200">Aop:200 - Estrogen receptor activation leading to breast cancer </a></td>
<td>MolecularInitiatingEvent</td>
</tr>
<tr>
<td><a href="/aops/167">Aop:167 - Early-life estrogen receptor agonism leading to endometrial adenosquamous carcinoma via promotion of sine oculis homeobox 1 progenitor cells</a></td>
<p><a name="_Hlk165971069">Taxonomic applicability:</a> In mammals there are two ER subtypes, ER alpha (ERα) and ER beta (ERβ), which are located on chromosome 6 and 14 and encoded by two different genes (ESR1 and ESR2) <a name="_Hlk162433655"></a>(Ascenzi et al., 2006). ERs were conventionally identified as mammal specific, but most vertebrates contain functional ERs. However, although teleost fish have receptors homologous to mammilian ERα, ERβ is divided into ERβ1 and ERβ2 resulting in three distinct ERs (Asnake et al., 2019; Menuet et al., 2004; Menuet et al., 2002). The majority of invertebrates (i.e. mollusks) possess a gene that is the orthologue of the vertebrate ER but in many species it has been demonstrated to only have constitutive transcriptional activity, and is not activated by ligand binding (Balbi et al., 2019). However, ERs in annelids share functional characteristics with vertebrate ERs and its transcriptional activity can be disrupted by known endocrine-disrupting substances (Keay & Thornton, 2009).</p>
<p>This event would generally be viewed as relevant to vertebrates, but not invertebrates.</p>
<p><a name="_Hlk165905099">Life stage:</a><a name="_Hlk165899451"> </a>This event is applicable to all life stages.</p>
<p>Sex: This event is applicable to both sexes.</p>
<h4>Key Event Description</h4>
<p>Site of action: The molecular site of action is the estrogen receptor (ER). ERs <a name="_Hlk162335835">are members of the steroid hormone receptor family which belongs to a group of nuclear receptors </a>that are transcriptionally activated by ligands leading to downstream activation of many cellular processes. ERs are composed of three principal domains – N-terminal domain (NTD), DNA binding domain (DBD), and the ligand binding domain (LBD). ER binds to specific DNA sequences known as estrogen response elements (EREs); EREs are generally short sequences located in the promoter region but can also exist in introns or exons (Klinge, 2001). ER-mediated gene transcription is initiated by binding of the DBD to an ERE with two distinct transcriptional activation domains, AF1 and AF2, located on the NTD and LBD respectively (Kumar et al., 2011).</p>
<p>Responses at the macromolecular level: ER’s bind to endogenous and exogenous compounds and are activated by endogenous ligands such as estrone (E1), estradiol (E2) and estriol (E3) (Ng et al., 2014). There are numerous compounds (e.g., natural or pharmaceutical estrogens, alkylphenols, organochlorine pesticides, phthalates, etc.) that can act as estrogen agonists or antagonists, and effectively mimic or block the natural effects of estrogens on the ER (Pillon et al., 2005; Schmieder et al., 2014).</p>
<p>ER is part of a multi-protein complex consisting of HSP 90, HSP 70, and immunophilins (Stice & Knowlton, 2008). In this multi-protein complex HSP 90 is the dominant protein and its binding to ER is essential for ER conformational binding of 17β-estradiol (Segnitz & Gehring, 1997). When binding on the LBD receptor occurs ER dissociates from HSP 90 and leads to receptor dimerization which can either be homodimers from the same isoform (ERα-Erα) or heterodimers containing one unit from both isoforms (ERα-Erβ) (Fliss et al., 2000). The translocation of these dimers into the nucleus modulates gene transcription (Aranda & Pascual, 2001).</p>
<h4>How it is Measured or Detected</h4>
<ul>
<li>OECD Test No. 455: Performance-based test guideline for stably transfected transactivation in vitro assays to detect estrogen receptor agonists and antagonists (OECD 2021).</li>
<li>OECD Test No. 457: BG1Luc Estrogen Receptor Transactivation Test Method for Identifying Estrogen Receptor Agonists and Antagonists (OECD 2012).</li>
<li>Standard Evaluation Procedure (SEP) for estrogen receptor transcriptional activation (Human Cell Line HeLa-9903) assay was developed by the U.S. Environmental Protection Agency (EPA).</li>
<li>ER-based transactivation assays that have been used to detect ER agonists and antagonist using cell lines include T47D-Kbluc assay (Wehmas et al., 2011), the ERα CALUX assay (Van et al.); MELN assay (Berckmans et al., 2007); and the yeast estrogen screen (YES; (De Boever et al., 2001)). The T47D-Kbluc assay responds to both ERα and ERß agonists but support the assumption that ERα is inducing more reporter expression than ERß. Each of these assays have undergone some level of validation.</li>
<li>Browne et al. (2015) integrated 18 ER ToxCast high-throughput screening (HTS) assays, measuring ER binding, dimerization, chromatin binding, transcriptional activation and ER-dependent cell proliferation, into the ToxCast ER pathway model. This mathematical model that in vitro assays to predict whether a chemical is an ER agonist or antagonist.</li>
<li>OECD Test No. 440: Uterotrophic Bioassay in Rodents: A Short-Term Screenign Test for Oestrogenic Properties. OCED Publishing. 2018. has been used to detect in vivo estrogenic activity.</li>
</ul>
<!-- end event text -->
</div>
<h4>References</h4>
<p>Aranda, A., & Pascual, A. (2001). Nuclear hormone receptors and gene expression. Physiological reviews, 81(3), 1269-1304.</p>
<p>Ascenzi, P., Bocedi, A., & Marino, M. (2006). Structure–function relationship of estrogen receptor α and β: Impact on human health. Molecular aspects of medicine, 27(4), 299-402.</p>
<p>Asnake, S., Modig, C., & Olsson, P.-E. (2019). Species differences in ligand interaction and activation of estrogen receptors in fish and human. The Journal of steroid biochemistry and molecular biology, 195, 105450.</p>
<p>Balbi, T., Ciacci, C., & Canesi, L. (2019). Estrogenic compounds as exogenous modulators of physiological functions in molluscs: Signaling pathways and biological responses. Comparative Biochemistry and Physiology Part C: Toxicology & Pharmacology, 222, 135-144.</p>
<p>Berckmans, P., Leppens, H., Vangenechten, C., & Witters, H. (2007). Screening of endocrine disrupting chemicals with MELN cells, an ER-transactivation assay combined with cytotoxicity assessment. Toxicology in vitro, 21(7), 1262-1267.</p>
<p>Browne, P., Judson, R. S., Casey, W. M., Kleinstreuer, N. C., & Thomas, R. S. (2015). Screening Chemicals for Estrogen Receptor Bioactivity Using a Computational Model. Environmental Science & Technology, 49(14), 8804-8814. <a href="https://doi.org/10.1021/acs.est.5b02641">https://doi.org/10.1021/acs.est.5b02641</a></p>
<p>De Boever, P., Demaré, W., Vanderperren, E., Cooreman, K., Bossier, P., & Verstraete, W. (2001). Optimization of a yeast estrogen screen and its applicability to study the release of estrogenic isoflavones from a soygerm powder. Environmental Health Perspectives, 109(7), 691-697.</p>
<p>Fliss, A. E., Benzeno, S., Rao, J., & Caplan, A. J. (2000). Control of estrogen receptor ligand binding by Hsp90. The Journal of steroid biochemistry and molecular biology, 72(5), 223-230.</p>
<p>Keay, J., & Thornton, J. W. (2009). Hormone-activated estrogen receptors in annelid invertebrates: implications for evolution and endocrine disruption. Endocrinology, 150(4), 1731-1738.</p>
<p>Klinge, C. M. (2001). Estrogen receptor interaction with estrogen response elements. Nucleic Acids Res, 29(14), 2905-2919. <a href="https://doi.org/10.1093/nar/29.14.2905">https://doi.org/10.1093/nar/29.14.2905</a></p>
<p>Kumar, R., Zakharov, M. N., Khan, S. H., Miki, R., Jang, H., Toraldo, G., Singh, R., Bhasin, S., & Jasuja, R. (2011). The dynamic structure of the estrogen receptor. Journal of amino acids, 2011.</p>
<p>Menuet, A., Le Page, Y., Torres, O., Kern, L., Kah, O., & Pakdel, F. (2004). Analysis of the estrogen regulation of the zebrafish estrogen receptor (ER) reveals distinct effects of ERalpha, ERbeta1 and ERbeta2. Journal of Molecular Endocrinology, 32(3), 975-986.</p>
<p>Menuet, A., Pellegrini, E., Anglade, I., Blaise, O., Laudet, V., Kah, O., & Pakdel, F. (2002). Molecular characterization of three estrogen receptor forms in zebrafish: binding characteristics, transactivation properties, and tissue distributions. Biology of reproduction, 66(6), 1881-1892.</p>
<p>Ng, H. W., Perkins, R., Tong, W., & Hong, H. (2014). Versatility or Promiscuity: The Estrogen Receptors, Control of Ligand Selectivity and an Update on Subtype Selective Ligands. International Journal of Environmental Research and Public Health, 11(9), 8709-8742. <a href="https://www.mdpi.com/1660-4601/11/9/8709">https://www.mdpi.com/1660-4601/11/9/8709</a></p>
<p>Pillon, A., Boussioux, A.-M., Escande, A., Aït-Aïssa, S., Gomez, E., Fenet, H., Ruff, M., Moras, D., Vignon, F., & Duchesne, M.-J. (2005). Binding of estrogenic compounds to recombinant estrogen receptor-α: application to environmental analysis. Environmental Health Perspectives, 113(3), 278-284.</p>
<p>Schmieder, P. K., Kolanczyk, R. C., Hornung, M. W., Tapper, M. A., Denny, J. S., Sheedy, B. R., & Aladjov, H. (2014). A rule-based expert system for chemical prioritization using effects-based chemical categories. SAR and QSAR in Environmental Research, 25(4), 253-287. <a href="https://doi.org/10.1080/1062936X.2014.898691">https://doi.org/10.1080/1062936X.2014.898691</a></p>
<p>Segnitz, B., & Gehring, U. (1997). The function of steroid hormone receptors is inhibited by the hsp90-specific compound geldanamycin. Journal of Biological Chemistry, 272(30), 18694-18701.</p>
<p>Stice, J. P., & Knowlton, A. A. (2008). Estrogen, NFκB, and the heat shock response. Molecular Medicine, 14, 517-527.</p>
<p>Van, d., Winter, R., Weimer, M., Beckmanns, P., Suzuki, G., Gijsberg, L., Jonas, A., Van, d. W., Hilda, & Aarts, J. Optimization and Prevalidation of the in Vitro ER CALUX Method to Test Estrogenic and Antiestrogenic Activity of Compounds.</p>
<p>Wehmas, L. C., Cavallin, J. E., Durhan, E. J., Kahl, M. D., Martinovic, D., Mayasich, J., Tuominen, T., Villeneuve, D. L., & Ankley, G. T. (2011). Screening complex effluents for estrogenic activity with the T47D‐KBluc cell bioassay: Assay optimization and comparison with in vivo responses in fish. Environmental toxicology and chemistry, 30(2), 439-445.</p>
<h3>List of Key Events in the AOP</h3>
<div>
<div>
<h4><a href="/events/78">Event: 78: Reduction, Cumulative fecundity and spawning</a><br></h4>
<h5>Short Name: Reduction, Cumulative fecundity and spawning</h5>
<p><p>Reduction of cumulative fecundity and spawning following exposure to low levels of TDCIPP (15, 46 and 90 nM) has been reported in 3 different zebrafish studies (Liu et al., 2013; Wang et al., 2015a; Zhu et al., 2015).</p>
</p>
<br>
<!-- end Evidence for Perturbation of This Event by Stressors -->
<h4>Domain of Applicability</h4>
<br>
<!-- loop to find taxonomic applicability under event -->
<p>Cumulative fecundity and spawning can, in theory, be evaluated for any egg laying animal.</p>
<br>
</div>
<!-- event text -->
<h4>Key Event Description</h4>
<p>Spawning refers to the release of eggs. Cumulative fecundity refers to the total number of eggs deposited by a female, or group of females over a specified period of time.</p>
<p>Cumulative fecundity and spawning can, in theory, be evaluated for any egg laying animal.</p>
<br>
<h4>Key Event Description</h4>
<p>Spawning refers to the release of eggs. Cumulative fecundity refers to the total number of eggs deposited by a female, or group of females over a specified period of time.</p>
<h4>How it is Measured or Detected</h4>
<p>In laboratory-based reproduction assays (e.g., OECD Test No. 229; OECD Test No. 240), spawning and cumulative fecundity can be directly measured through daily observation of egg deposition and egg counts.</p>
<h4>How it is Measured or Detected</h4>
<p>In laboratory-based reproduction assays (e.g., OECD Test No. 229; OECD Test No. 240), spawning and cumulative fecundity can be directly measured through daily observation of egg deposition and egg counts.</p>
<p>In some cases, fecundity may be estimated based on gonado-somatic index (<a href="http://www.oecd.org/officialdocuments/publicdisplaydocumentpdf/?cote=env/jm/mono(2008)22&doclanguage=en">OECD 2008</a>).</p>
<br>
<h4>Regulatory Significance of the AO</h4>
<p>Cumulative fecundity is the most apical endpoint considered in the OECD 229 Fish Short Term Reproduction Assay. The OECD 229 assay serves as screening assay for endocrine disruption and associated reproductive impairment (<a href="http://www.oecd-ilibrary.org/environment/test-no-229-fish-short-term-reproduction-assay_9789264185265-en">OECD 2012</a>). Fecundity is also an important apical endpoint in the Medaka Extended One Generation Reproduction Test (MEOGRT; <a href="http://www.oecd-ilibrary.org/environment/test-no-240-medaka-extended-one-generation-reproduction-test-meogrt_9789264242258-en">OECD Test Guideline 240</a>; OECD 2015).</p>
<h4>Regulatory Significance of the AO</h4>
<p>Cumulative fecundity is the most apical endpoint considered in the OECD 229 Fish Short Term Reproduction Assay. The OECD 229 assay serves as screening assay for endocrine disruption and associated reproductive impairment (<a href="http://www.oecd-ilibrary.org/environment/test-no-229-fish-short-term-reproduction-assay_9789264185265-en">OECD 2012</a>). Fecundity is also an important apical endpoint in the Medaka Extended One Generation Reproduction Test (MEOGRT; <a href="http://www.oecd-ilibrary.org/environment/test-no-240-medaka-extended-one-generation-reproduction-test-meogrt_9789264242258-en">OECD Test Guideline 240</a>; OECD 2015).</p>
<p>A variety of fish life cycle tests also include cumulative fecundity as an endpoint (<a href="http://www.oecd.org/officialdocuments/publicdisplaydocumentpdf/?cote=env/jm/mono(2008)22&doclanguage=en">OECD 2008</a>).</p>
<p> </p>
<br>
<h4>References</h4>
<ul>
<h4>References</h4>
<ul>
<li>OECD 2008. Series on testing and assessment, Number 95. Detailed Review Paper on Fish Life-cycle Tests. OECD Publishing, Paris. ENV/JM/MONO(2008)22.</li>
<li>OECD (2015), <em>Test No. 240: Medaka Extended One Generation Reproduction Test (MEOGRT)</em>, OECD Publishing, Paris.<br />
<!-- Evidence for Perturbation of This Event by Stressors -->
<!-- end Evidence for Perturbation of This Event by Stressors -->
<!-- event text -->
<h4>Domain of Applicability</h4>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt"><strong>Taxonomic applicability:</strong> </span><span style="font-size:11pt">Oviparous vertebrates synthesize yolk precursor proteins that are transported in the circulation for uptake by developing oocytes. Many invertebrates also synthesize vitellogenins that are taken up into developing oocytes via active transport mechanisms. However, invertebrate vitellogenins are transported in hemolymph or via other transport mechanisms rather than plasma.</span></span></span></p>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt"><strong>Life stage: </strong>This KE is applicable to all life stages following the differentiation of the liver. Embryos prior to liver differentiation should not be included.</span></span></span></p>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt"><strong>Sex: </strong>This KE is applicable to both sexes.</span></span></span></p>
<h4>Key Event Description</h4>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Vitellogenins are large serum phospholipoglycoprotein that are encoded by a family of paralog genes whose number varies in the different vertebrate lineages resulting in numerous isoforms (Wahli, 1988). Vtg is synthesized in the liver and is secreted into the blood as ~500 kDa homodimers which circulate to the ovaries for uptake and bind to receptors on the surface of growing oocytes (Wallace, 1985).</span></span></span></p>
<!-- end event text -->
</div>
<h4>How it is Measured or Detected</h4>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Vitellogenin concentrations in plasma are typically measured using enzyme linked immunosorbent assay (ELISA; e.g., Denslow et al., 1999; Holbech et al., 2001). Less specific and/or sensitive assays such as determination of alkali-labile phosphoprotein (e.g., Kramer et al., 1998) and Western blotting (e.g., Heppell et al., 1995) may also be used.</span></span></span></p>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">There are also several standardized test guidelines that measure vtg including: Fish Short Term Reproduction Assay (OECD, 2009a), 21-day Fish Assay (OECD, 2009b); Fish Sexual Development Test (OECD, 2011), Medaka Extended One Generation Reproduction Test (OECD, 2015a). Measurement of vtg is also an optional parameter in the Larval Amphibian Growth and Development Assay (OECD, 2015b). The US Environmental Protection Agency (EPA) has similar standardized guidelines (US EPA, 2009, US EPA, 2014) as does the EU as part of the Guidance For The Identification Of Endocrine Disruptors In The Context Of Regulations (EC 2013, EC 2018).</span></span></span></p>
<div>
<div>
<h4><a href="/events/307">Event: 307: Increase, Vitellogenin synthesis in liver</a><br></h4>
<h5>Short Name: Increase, Vitellogenin synthesis in liver</h5>
</div>
<h4>References</h4>
<ul>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Denslow, N. D., Chow, M. C., Kroll, K. J., & Green, L. (1999). Vitellogenin as a biomarker of exposure for estrogen or estrogen mimics. <em>Ecotoxicology</em>,<em> 8</em>, 385-398. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Heppell, S. A., Denslow, N. D., Folmar, L. C., & Sullivan, C. V. (1995). Universal assay of vitellogenin as a biomarker for environmental estrogens. <em>Environmental Health Perspectives</em>,<em> 103</em>(suppl 7), 9-15. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Holbech, H., Andersen, L., Petersen, G. I., Korsgaard, B., Pedersen, K. L., & Bjerregaard, P. (2001). Development of an ELISA for vitellogenin in whole body homogenate of zebrafish (Danio rerio). <em>Comparative Biochemistry and Physiology Part C: Toxicology & Pharmacology</em>,<em> 130</em>(1), 119-131. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Kramer, V., Miles-Richardson, S., Pierens, S., & Giesy, J. (1998). Reproductive impairment and induction of alkaline-labile phosphate, a biomarker of estrogen exposure, in fathead minnows (Pimephales promelas) exposed to waterborne 17β-estradiol. <em>Aquatic Toxicology</em>,<em> 40</em>(4), 335-360. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Wahli, W. (1988). Evolution and expression of vitellogenin genes. <em>Trends in Genetics</em>,<em> 4</em>(8), 227-232. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Wallace, R. A. (1985). Vitellogenesis and oocyte growth in nonmammalian vertebrates. <em>Oogenesis</em>, 127-177. </span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Although vitellogenin is conserved among oviparous vertebrates and many invertebrates, liver is not a relevant tissue for the production of vitellogenin in invertebrates (Wahli, 1988).</span></span></span></li>
</ul>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt"><strong>Life stage: </strong>This KE is applicable to all life stages following the differentiation of the liver. Embryos prior to liver differentiation should not be included.</span></span></span></p>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt"><strong>Sex: </strong>This KE is applicable to both sexes.</span></span></span></p>
<h4>Key Event Description</h4>
<p><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt"><a name="_Hlk167176446"><span style="color:#333399">Vitellogenin (VTG) is an egg yolk precursor protein synthesized by hepatocytes of oviparous vertebrates </span></a><span style="color:#333399">(Hara et al., 2016). Transcription of <em>vtg</em> is regulated by estrogens and their interaction on ERs. </span></span></span><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">In males expression can be modulated by exogenous compounds. Under high estrogen stimulation the fold increase of <em>vtg</em> transcripts increases by orders of magnitude (Brock & Shapiro, 1983).</span></span></span></p>
<!-- end event text -->
</div>
<h4>How it is Measured or Detected</h4>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Relative abundance of vitellogenin transcripts or protein can be measured in liver tissue (e.g., Miracle et al., 2006), hepatocytes (e.g., Vaillant et al., 1988), exposed in vitro, or whole-body homogenates from organisms exposed in vivo (Holbech et al., 2001). </span></span></span></p>
<p><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt"><span style="color:#333399">mRNA transcripts can be measured using real-time quantitative polymerase chain reaction (qPCR) while protein quantification can be measured using alkali-labile phosphoprotein (e.g., Kramer et al., 1998), or immunochemical methods such as radioimmunoassay (RIA; e.g., Tyler & Sumpter, 1990), enzyme linked immunosorbent </span><a name="_Hlk165465390"><span style="color:#333399">assay </span></a><span style="color:#333399">(ELISA; e.g., Denslow et al., 1999), and Western blotting (e.g., Heppell et al., 1995).</span></span></span></p>
<div>
<div>
<h4><a href="/events/252">Event: 252: Increase, Renal pathology due to VTG deposition</a><br></h4>
<h5>Short Name: Increase, Renal pathology due to VTG deposition</h5>
</div>
<h4>References</h4>
<ul>
<li><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Brock, M. L., & Shapiro, D. (1983). Estrogen regulates the absolute rate of transcription of the Xenopus laevis vitellogenin genes. <em>Journal of Biological Chemistry</em>,<em> 258</em>(9), 5449-5455. </span></span></li>
<li><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Denslow, N. D., Chow, M. C., Kroll, K. J., & Green, L. (1999). Vitellogenin as a biomarker of exposure for estrogen or estrogen mimics. <em>Ecotoxicology</em>,<em> 8</em>, 385-398. </span></span></li>
<li><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Hara, A., Hiramatsu, N., & Fujita, T. (2016). Vitellogenesis and choriogenesis in fishes. <em>Fisheries Science</em>,<em> 82</em>(2), 187-202. <a href="https://doi.org/10.1007/s12562-015-0957-5" style="color:#0563c1; text-decoration:underline">https://doi.org/10.1007/s12562-015-0957-5</a> </span></span></li>
<li><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Heppell, S. A., Denslow, N. D., Folmar, L. C., & Sullivan, C. V. (1995). Universal assay of vitellogenin as a biomarker for environmental estrogens. <em>Environmental Health Perspectives</em>,<em> 103</em>(suppl 7), 9-15. </span></span></li>
<li><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Holbech, H., Andersen, L., Petersen, G. I., Korsgaard, B., Pedersen, K. L., & Bjerregaard, P. (2001). Development of an ELISA for vitellogenin in whole body homogenate of zebrafish (Danio rerio). <em>Comparative Biochemistry and Physiology Part C: Toxicology & Pharmacology</em>,<em> 130</em>(1), 119-131. </span></span></li>
<li><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Kramer, V., Miles-Richardson, S., Pierens, S., & Giesy, J. (1998). Reproductive impairment and induction of alkaline-labile phosphate, a biomarker of estrogen exposure, in fathead minnows (Pimephales promelas) exposed to waterborne 17β-estradiol. <em>Aquatic Toxicology</em>,<em> 40</em>(4), 335-360. </span></span></li>
<li><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Miracle, A., Ankley, G., & Lattier, D. (2006). Expression of two vitellogenin genes (vg1 and vg3) in fathead minnow (Pimephales promelas) liver in response to exposure to steroidal estrogens and androgens. <em>Ecotoxicology and environmental safety</em>,<em> 63</em>(3), 337-342. </span></span></li>
<li><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Tyler, C. R., & Sumpter, J. P. (1990). The development of a radioimmunoassay for carp, Cyprinus carpio, vitellogenin. <em>Fish Physiology and Biochemistry</em>,<em> 8</em>, 129-140. </span></span></li>
<li><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Vaillant, C., Le Guellec, C., Pakdel, F., & Valotaire, Y. (1988). Vitellogenin gene expression in primary culture of male rainbow trout hepatocytes. <em>General and Comparative Endocrinology</em>,<em> 70</em>(2), 284-290. </span></span></li>
<li><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Wahli, W. (1988). Evolution and expression of vitellogenin genes. <em>Trends in Genetics</em>,<em> 4</em>(8), 227-232. </span></span></li>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt"><strong>Life stage: </strong>This KE is applicable to all life stages following the differentiation of the kidney.</span></span></span></p>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt"><strong>Sex: </strong>This KE is applicable to both sexes.</span></span></span></p>
<h4>Key Event Description</h4>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Renal pathology deals with the characterization of the kidneys. The kidneys perform a suite of physiological roles that are critical for organismal homeostasis including waste excretion, osmoregulation, and fluid homeostasis (Preuss, 1993). Each kidney is made up of specialized epithelial cells known as nephrons and while nephron numbers can vary greatly between species their overall function remains conserved in vertebrates (Desgrange & Cereghini, 2015). Nephrons act as filtering units that are composed of glomeruli and tubules which are responsible for removing metabolic waste from the bloodstream, regulating fluids, and balancing electrolytes (Wesselman et al., 2023). Organ tissue damage can occur after exposure to toxins, parasites, or be caused by disease. If pathology is measurable this would be an indication of damage or diseased tissue state and a departure from normal/healthy tissue.</span></span></span></p>
<h4>How it is Measured or Detected</h4>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Histopathology focuses on the changes in tissues and is a technique used for identifying correlations with biochemical markers. Generally renal pathology is measured after either whole organism or specific tissue of interest is fixed, dehydrated, and then embedded in wax, commonly paraffin wax. Sections are then cut to approximately 3–5 μm in thickness and stained before being examined under a microscope (e.g., Folmar et al., 2001; Mihaich et al., 2012; Zha et al., 2007).</span></span></span></p>
<ul>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">OECD Test No. 123: Guidance document on the diagnosis of endocrine-related histopathology in fish gonads (OECD 2010).</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">OECD Test No. 227: Guidance document on medaka histopathology techniques and evaluation for the medaka extended one-generation reproduction test (OECD 2015)</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Crissman et al. (2004) describes best practice guidelines for toxicologic histopathology.</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Fiedler et al. (2023) have written standardized tissue sampling guidelines for histopathological analyses using rainbow trout.</span></span></span></li>
</ul>
<!-- end event text -->
</div>
<h4>References</h4>
<ul>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Crissman, J. W., Goodman, D. G., Hildebrandt, P. K., Maronpot, R. R., Prater, D. A., Riley, J. H., Seaman, W. J., & Thake, D. C. (2004). Best Practices Guideline: Toxicologic Histopathology. <em>Toxicologic Pathology</em>,<em> 32</em>(1), 126-131. <a href="https://doi.org/10.1080/01926230490268756" style="color:#0563c1; text-decoration:underline">https://doi.org/10.1080/01926230490268756</a> </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Desgrange, A., & Cereghini, S. (2015). Nephron patterning: lessons from Xenopus, zebrafish, and mouse studies. <em>Cells</em>,<em> 4</em>(3), 483-499. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Fiedler, S., Schrader, H., Theobalt, N., Hofmann, I., Geiger, T., Arndt, D., Wanke, R., Schwaiger, J., & Blutke, A. (2023). Standardized tissue sampling guidelines for histopathological and molecular analyses of rainbow trout (Oncorhynchus mykiss) in ecotoxicological studies. <em>PLOS ONE</em>,<em> 18</em>(7), e0288542. <a href="https://doi.org/10.1371/journal.pone.0288542" style="color:#0563c1; text-decoration:underline">https://doi.org/10.1371/journal.pone.0288542</a> </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Folmar, L. C., Gardner, G. R., Schreibman, M. P., Magliulo-Cepriano, L., Mills, L. J., Zaroogian, G., Gutjahr-Gobell, R., Haebler, R., Horowitz, D. B., & Denslow, N. D. (2001). Vitellogenin-induced pathology in male summer flounder (Paralichthys dentatus). <em>Aquatic Toxicology</em>,<em> 51</em>(4), 431-441. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Mihaich, E., Rhodes, J., Wolf, J., van der Hoeven, N., Dietrich, D., Hall, A. T., Caspers, N., Ortego, L., Staples, C., & Dimond, S. (2012). Adult fathead minnow, Pimephales promelas, partial life‐cycle reproductive and gonadal histopathology study with bisphenol A. <em>Environmental toxicology and chemistry</em>,<em> 31</em>(11), 2525-2535. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Preuss, H. G. (1993). Basics of renal anatomy and physiology. <em>Clinics in laboratory medicine</em>,<em> 13</em>(1), 1-11. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Wesselman, H. M., Gatz, A. E., Pfaff, M. R., Arceri, L., & Wingert, R. A. (2023). Estrogen signaling influences nephron segmentation of the zebrafish embryonic kidney. <em>Cells</em>,<em> 12</em>(4), 666. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Zha, J., Wang, Z., Wang, N., & Ingersoll, C. (2007). Histological alternation and vitellogenin induction in adult rare minnow (Gobiocypris rarus) after exposure to ethynylestradiol and nonylphenol. <em>Chemosphere</em>,<em> 66</em>(3), 488-495. </span></span></li>
</ul>
<h3>List of Adverse Outcomes in this AOP</h3>
<div>
<div>
<h4><a href="/events/360">Event: 360: Decrease, Population trajectory</a><br></h4>
<h5>Short Name: Decrease, Population trajectory</h5>
<td><a href="/aops/23">Aop:23 - Androgen receptor agonism leading to reproductive dysfunction (in repeat-spawning fish)</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/25">Aop:25 - Aromatase inhibition leading to reproductive dysfunction</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/29">Aop:29 - Estrogen receptor agonism leading to reproductive dysfunction</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/30">Aop:30 - Estrogen receptor antagonism leading to reproductive dysfunction</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/100">Aop:100 - Cyclooxygenase inhibition leading to reproductive dysfunction via inhibition of female spawning behavior</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/122">Aop:122 - Prolyl hydroxylase inhibition leading to reproductive dysfunction via increased HIF1 heterodimer formation</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/123">Aop:123 - Unknown MIE leading to reproductive dysfunction via increased HIF-1alpha transcription</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/155">Aop:155 - Deiodinase 2 inhibition leading to reduced young of year survival via posterior swim bladder inflation</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/156">Aop:156 - Deiodinase 2 inhibition leading to reduced young of year survival via anterior swim bladder inflation</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/157">Aop:157 - Deiodinase 1 inhibition leading to reduced young of year survival via posterior swim bladder inflation </a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/158">Aop:158 - Deiodinase 1 inhibition leading to reduced young of year survival via anterior swim bladder inflation</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/159">Aop:159 - Thyroperoxidase inhibition leading to reduced young of year survival via anterior swim bladder inflation</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/101">Aop:101 - Cyclooxygenase inhibition leading to reproductive dysfunction via inhibition of pheromone release</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/102">Aop:102 - Cyclooxygenase inhibition leading to reproductive dysfunction via interference with meiotic prophase I /metaphase I transition</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/63">Aop:63 - Cyclooxygenase inhibition leading to reproductive dysfunction</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/103">Aop:103 - Cyclooxygenase inhibition leading to reproductive dysfunction via interference with spindle assembly checkpoint</a></td>
<td><a href="/aops/23">Aop:23 - Androgen receptor agonism leading to reproductive dysfunction (in repeat-spawning fish)</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/25">Aop:25 - Aromatase inhibition leading to reproductive dysfunction</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/29">Aop:29 - Estrogen receptor agonism leading to reproductive dysfunction</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/30">Aop:30 - Estrogen receptor antagonism leading to reproductive dysfunction</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/100">Aop:100 - Cyclooxygenase inhibition leading to reproductive dysfunction via inhibition of female spawning behavior</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/122">Aop:122 - Prolyl hydroxylase inhibition leading to reproductive dysfunction via increased HIF1 heterodimer formation</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/123">Aop:123 - Unknown MIE leading to reproductive dysfunction via increased HIF-1alpha transcription</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/155">Aop:155 - Deiodinase 2 inhibition leading to increased mortality via reduced posterior swim bladder inflation</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/156">Aop:156 - Deiodinase 2 inhibition leading to increased mortality via reduced anterior swim bladder inflation</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/157">Aop:157 - Deiodinase 1 inhibition leading to increased mortality via reduced posterior swim bladder inflation </a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/158">Aop:158 - Deiodinase 1 inhibition leading to increased mortality via reduced anterior swim bladder inflation</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/159">Aop:159 - Thyroperoxidase inhibition leading to increased mortality via reduced anterior swim bladder inflation</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/101">Aop:101 - Cyclooxygenase inhibition leading to reproductive dysfunction via inhibition of pheromone release</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/102">Aop:102 - Cyclooxygenase inhibition leading to reproductive dysfunction via interference with meiotic prophase I /metaphase I transition</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/63">Aop:63 - Cyclooxygenase inhibition leading to reproductive dysfunction</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/103">Aop:103 - Cyclooxygenase inhibition leading to reproductive dysfunction via interference with spindle assembly checkpoint</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/292">Aop:292 - Inhibition of tyrosinase leads to decreased population in fish</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/310">Aop:310 - Embryonic Activation of the AHR leading to Reproductive failure, via epigenetic down-regulation of GnRHR</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/16">Aop:16 - Acetylcholinesterase inhibition leading to acute mortality</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/312">Aop:312 - Acetylcholinesterase Inhibition leading to Acute Mortality via Impaired Coordination & Movement</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/334">Aop:334 - Glucocorticoid Receptor Agonism Leading to Impaired Fin Regeneration</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/336">Aop:336 - DNA methyltransferase inhibition leading to population decline (1)</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/337">Aop:337 - DNA methyltransferase inhibition leading to population decline (2)</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/338">Aop:338 - DNA methyltransferase inhibition leading to population decline (3)</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/339">Aop:339 - DNA methyltransferase inhibition leading to population decline (4)</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/340">Aop:340 - DNA methyltransferase inhibition leading to transgenerational effects (1)</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/341">Aop:341 - DNA methyltransferase inhibition leading to transgenerational effects (2)</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/289">Aop:289 - Inhibition of 5α-reductase leading to impaired fecundity in female fish</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/297">Aop:297 - Inhibition of retinaldehyde dehydrogenase leads to population decline</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/346">Aop:346 - Aromatase inhibition leads to male-biased sex ratio via impacts on gonad differentiation</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/363">Aop:363 - Thyroperoxidase inhibition leading to altered visual function via altered retinal layer structure</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/349">Aop:349 - Inhibition of 11β-hydroxylase leading to decresed population trajectory </a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/348">Aop:348 - Inhibition of 11β-Hydroxysteroid Dehydrogenase leading to decreased population trajectory </a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/376">Aop:376 - Androgen receptor agonism leading to male-biased sex ratio</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/386">Aop:386 - Deposition of ionizing energy leading to population decline via inhibition of photosynthesis</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/387">Aop:387 - Deposition of ionising energy leading to population decline via mitochondrial dysfunction</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/388">Aop:388 - Deposition of ionising energy leading to population decline via programmed cell death</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/389">Aop:389 - Oxygen-evolving complex damage leading to population decline via inhibition of photosynthesis</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/364">Aop:364 - Thyroperoxidase inhibition leading to altered visual function via decreased eye size</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/365">Aop:365 - Thyroperoxidase inhibition leading to altered visual function via altered photoreceptor patterning</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/399">Aop:399 - Inhibition of Fyna leading to increased mortality via decreased eye size (Microphthalmos)</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/410">Aop:410 - GSK3beta inactivation leading to increased mortality via defects in developing inner ear</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/216">Aop:216 - Deposition of energy leading to population decline via DNA strand breaks and follicular atresia</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/238">Aop:238 - Deposition of energy leading to population decline via DNA strand breaks and oocyte apoptosis</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/299">Aop:299 - Deposition of energy leading to population decline via DNA oxidation and follicular atresia</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/311">Aop:311 - Deposition of energy leading to population decline via DNA oxidation and oocyte apoptosis</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/444">Aop:444 - Ionizing radiation leads to reduced reproduction in Eisenia fetida via reduced spermatogenesis and cocoon hatchability</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/138">Aop:138 - Organic anion transporter (OAT1) inhibition leading to renal failure and mortality</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/177">Aop:177 - Cyclooxygenase 1 (COX1) inhibition leading to renal failure and mortality</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/97">Aop:97 - 5-hydroxytryptamine transporter (5-HTT; SERT) inhibition leading to population decline</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/203">Aop:203 - 5-hydroxytryptamine transporter inhibition leading to decreased reproductive success and population decline</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/218">Aop:218 - Inhibition of CYP7B activity leads to decreased reproductive success via decreased locomotor activity</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/219">Aop:219 - Inhibition of CYP7B activity leads to decreased reproductive success via decreased sexual behavior</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/323">Aop:323 - PPARalpha Agonism Leading to Decreased Viable Offspring via Decreased 11-Ketotestosterone</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/536">Aop:536 - Estrogen receptor agonism leading to reduced survival and population growth due to renal failure</a></td>
<td>KeyEvent</td>
</tr>
<tr>
<td><a href="/aops/540">Aop:540 - Oxidative Stress in the Fish Ovary Leads to Reproductive Impairment via Reduced Vitellogenin Production</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/564">Aop:564 - DBDPE-induced inhibition of mitochondrial complex Ⅰ leading to population decline via neurotoxicity and metabotoxicity.</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/567">Aop:567 - Binding of plastoquinone B leading to decrease, population growth rate via decrease, photosystem II efficiency</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/592">Aop:592 - DBDPE-induced DNA strand breaks and LDH activity inhibition leading to population growth rate decline via energy metabolism disrupt and apoptosis</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/605">Aop:605 - Thyroid Peroxidase Inhibition Leading to Reduced, Swimming Performance via Hypomyelination</a></td>
<td>AdverseOutcome</td>
</tr>
<tr>
<td><a href="/aops/608">Aop:608 - Thyroid Hormone Excess Leading to Reduced, Swimming Performance via Hypomyelination</a></td>
<p>Consideration of population size and changes in population size over time is potentially relevant to all living organisms.</p>
<br>
</div>
<!-- event text -->
<h4>Key Event Description</h4>
<p>Maintenance of sustainable fish and wildlife populations (i.e., adequate to ensure long-term delivery of valued ecosystem services) is an accepted regulatory goal upon which risk assessments and risk management decisions are based.</p>
<br>
<p>Consideration of population size and changes in population size over time is potentially relevant to all living organisms.</p>
<h4>How it is Measured or Detected</h4>
<p>Population trajectories, either hypothetical or site specific, can be estimated via population modeling based on measurements of vital rates or reasonable surrogates measured in laboratory studies. As an example, Miller and Ankley 2004 used measures of cumulative fecundity from laboratory studies with repeat spawning fish species to predict population-level consequences of continuous exposure.</p>
<br>
<h4>Key Event Description</h4>
<p style="text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">A population can be defined as a group of interbreeding organisms, all of the same species, occupying a specific space during a specific time (Vandermeer and Goldberg 2003, Gotelli 2008). As the population is the biological level of organization that is often the focus of ecological risk</span> <span style="color:black">assessments, population growth rate (and hence population size over time) is important to consider within the context of applied conservation practices.</span></span></span></span></p>
<p style="text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">If N is the size of the population and t is time, then the population growth rate (dN/dt) is proportional to the instantaneous rate of increase, r, which measures the per capita rate of population increase over a short time interval. Therefore, r, is a difference between the instantaneous birth rate (number of births per individual per unit of time; b) and the instantaneous death rate (number of deaths per individual per unit of time; d) [Equation 1]. Because r is an instantaneous rate, its units can be changed via division. For example, as there are 24 hours in a day, an r of 24 individuals/(individual x day) is equal to an r of 1 individual/(individual/hour) (Caswell 2001, Vandermeer and Goldberg 2003, Gotelli 2008, Murray and Sandercock 2020). </span></span></span></span></p>
<p style="margin-left:144px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">Equation 1: r = b - d</span></span></span></span></p>
<p style="text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">This key event refers to scenarios where r < 0 (instantaneous death rate exceeds instantaneous birth rate).</span></span></span></span></p>
<p style="text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">Examining r in the context of population growth rate:</span></span></span></span></p>
<p style="margin-left:48px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">● A population will decrease to extinction when the instantaneous death rate exceeds the instantaneous birth rate (r < 0). </span></span></span></span></p>
<p style="text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black"> ● The smaller the value of r below 1, the faster the population will decrease to zero. </span></span></span></span></p>
<p style="margin-left:48px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">● A population will increase when resources are available and the instantaneous birth rate exceeds the instantaneous death rate (r > 0)</span></span></span></span></p>
<p style="text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black"> ● The larger the value that r exceeds 1, the faster the population can increase over time </span></span></span></span></p>
<p style="margin-left:48px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">● A population will neither increase or decrease when the population growth rate equals 0 (either due to N = 0, or if the per capita birth and death rates are exactly balanced). For example, the per capita birth and death rates could become exactly balanced due to density dependence and/or to the effect of a stressor that reduces survival and/or reproduction (Caswell 2001, Vandermeer and Goldberg 2003, Gotelli 2008, Murray and Sandercock 2020). </span></span></span></span></p>
<p style="text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">Effects incurred on a population from a chemical or non-chemical stressor could have an impact directly upon birth rate (reproduction) and/or death rate (survival), thereby causing a decline in population growth rate. </span></span></span></span></p>
<p style="margin-left:48px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">● Example of direct effect on r: Exposure to 17b-trenbolone reduced reproduction (i.e., reduced b) in the fathead minnow over 21 days at water concentrations ranging from 0.0015 to about 41 mg/L (Ankley et al. 2001; Miller and Ankley 2004). </span></span></span></span></p>
<p style="text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">Alternatively, a stressor could indirectly impact survival and/or reproduction. </span></span></span></span></p>
<p style="margin-left:48px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">● Example of indirect effect on r: Exposure of non-sexually differentiated early life stage fathead minnow to the fungicide prochloraz has been shown to produce male-biased sex ratios based on gonad differentiation, and resulted in projected change in population growth rate (decrease in reproduction due to a decrease in females and thus recruitment) using a population model. (Holbech et al., 2012; Miller et al. 2022)</span></span></span></span></p>
<p style="text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">Density dependence can be an important consideration:</span></span></span></span></p>
<p style="margin-left:48px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">● The effect of density dependence depends upon the quantity of resources present within a landscape. A change in available resources could increase or decrease the effect of density dependence and therefore cause a change in population growth rate via indirectly impacting survival and/or reproduction. </span></span></span></span></p>
<p style="margin-left:48px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">● This concept could be thought of in terms of community level interactions whereby one species is not impacted but a competitor species is impacted by a chemical stressor resulting in a greater availability of resources for the unimpacted species. In this scenario, the impacted species would experience a decline in population growth rate. The unimpacted species would experience an increase in population growth rate (due to a smaller density dependent effect upon population growth rate for that species). </span> </span></span></span></p>
<p style="text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">Closed versus open systems:</span></span></span></span></p>
<p style="margin-left:48px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">● The above discussion relates to closed systems (there is no movement of individuals between population sites) and thus a declining population growth rate cannot be augmented by immigration. </span></span></span></span></p>
<p style="margin-left:48px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">● When individuals depart (emigrate out of a population) the loss will diminish population growth rate. </span></span></span></span></p>
<p style="text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">Population growth rate applies to all organisms, both sexes, and all life stages.</span></span></span></span></p>
<p> </p>
<h4>Regulatory Significance of the AO</h4>
<p>Maintenance of sustainable fish and wildlife populations (i.e., adequate to ensure long-term delivery of valued ecosystem services) is a widely accepted regulatory goal upon which risk assessments and risk management decisions are based.</p>
<h4>How it is Measured or Detected</h4>
<p style="text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">Population growth rate (instantaneous growth rate) can be measured by sampling a population over an interval of time (i.e. from time t = 0 to time t = 1). The interval of time should be selected to correspond to the life history of the species of interest (i.e. will be different for rapidly growing versus slow growing populations). The population growth rate, r, can be determined by taking the difference (subtracting) between the initial population size, N</span><sub><span style="font-size:9pt"><span style="color:black">t=0 </span></span></sub><span style="color:black">(population size at time t=0), and the population size at the end of the interval, N</span><sub><span style="font-size:9pt"><span style="color:black">t=1 </span></span></sub><span style="color:black">(population size at time t = 1), and then subsequently dividing by the initial population size. </span></span></span></span></p>
<p style="text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">The diversity of forms, sizes, and life histories among species has led to the development of a vast number of field techniques for estimation of population size and thus population growth over time (Bookhout 1994, McComb et al. 2021). </span></span></span></span></p>
<p style="margin-left:48px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">● For stationary species an observational strategy may involve dividing a habitat into units. After setting up the units, samples are performed throughout the habitat at a select number of units (determined using a statistical sampling design) over a time interval (at time t = 0 and again at time t = 1), and the total number of organisms within each unit are counted. The numbers recorded are assumed to be representative for the habitat overall, and can be used to estimate the population growth rate within the entire habitat over the time interval. </span></span></span></span></p>
<p style="margin-left:48px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">● For species that are mobile throughout a large range, a strategy such as using a mark-recapture method may be employed (i.e. tags, bands, transmitters) to determine a count over a time interval (at time = 0 and again at time =1). </span></span></span></span></p>
<p style="text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">Population growth rate can also be estimated using mathematical model constructs (for example, ranging from simple differential equations to complex age or stage structured matrix projection models and individual based modeling approaches), and may assume a linear or nonlinear population increase over time (Caswell 2001, Vandermeer and Goldberg 2003, Gotelli 2008, Murray and Sandercock 2020). The AOP framework can be used to support the translation of pathway-specific mechanistic data into responses relevant to population models and output from the population models, such as changing (declining) population growth rate, can be used to assess and manage risks of chemicals (Kramer et al. 2011). As such, this translational capability can increase the capacity and efficiency of safety assessments both for single chemicals and chemical mixtures (Kramer et al. 2011). </span></span></span></span></p>
<p style="text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">Some examples of modeling constructs used to investigate population growth rate:</span></span></span></span></p>
<p style="margin-left:48px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">● A modeling construct could be based upon laboratory toxicity tests to determine effect(s) that are then linked to the population model and used to estimate decline in population growth rate. Miller et al. (2007) used concentration–response data from short term reproductive assays with fathead minnow (<em>Pimephales promelas</em>) exposed to endocrine disrupting chemicals in combination with a population model to examine projected alterations in population growth rate. </span></span></span></span></p>
<p style="margin-left:48px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">● A model construct could be based upon a combination of effects-based monitoring at field sites (informed by an AOP) and a population model. Miller et al. (2015) applied a population model informed by an AOP to project declines in population growth rate for white suckers (Catostomus commersoni) using observed changes in sex steroid synthesis in fish exposed to a complex pulp and paper mill effluent in Jackfish Bay, Ontario, Canada. Furthermore, a model construct could be comprised of a series of quantitative models using KERs that culminates in the estimation of change (decline) in population growth rate. </span></span></span></span></p>
<p style="margin-left:48px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">● A quantitative adverse outcome pathway (qAOP) has been defined as a mathematical construct that models the dose–response or response–response relationships of all KERs described in an AOP (Conolly et al. 2017, Perkins et al. 2019). Conolly et al. (2017) developed a qAOP using data generated with the aromatase inhibitor fadrozole as a stressor and then used it to predict potential population‐level impacts (including decline in population growth rate). The qAOP modeled aromatase inhibition (the molecular initiating event) leading to reproductive dysfunction in fathead minnow (Pimephales promelas) using 3 computational models: a hypothalamus–pituitary–gonadal axis model (based on ordinary differential equations) of aromatase inhibition leading to decreased vitellogenin production (Cheng et al. 2016), a stochastic model of oocyte growth dynamics relating vitellogenin levels to clutch size and spawning intervals (Watanabe et al. 2016), and a population model (Miller et al. 2007).</span></span></span></span></p>
<p style="margin-left:48px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">● Dynamic energy budget (DEB) models offer a methodology that reverse engineers stressor effects on growth, reproduction, and/or survival into modular characterizations related to the acquisition and processing of energy resources (Nisbet et al. 2000, Nisbet et al. 2011). Murphy et al. (2018) developed a conceptual model to link DEB and AOP models by interpreting AOP key events as measures of damage-inducing processes affecting DEB variables and rates.</span></span></span></span></p>
<p style="margin-left:48px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">● Endogenous Lifecycle Models (ELMs), capture the endogenous lifecycle processes of growth, development, survival, and reproduction and integrate these to estimate and predict expected fitness (Etterson and Ankley, 2021). AOPs can be used to inform ELMs of effects of chemical stressors on the vital rates that determine fitness, and to decide what hierarchical models of endogenous systems should be included within an ELM (Etterson and Ankley, 2021).</span></span></span></span></p>
<p> </p>
<br>
<h4>Regulatory Significance of the AO</h4>
<p>Maintenance of sustainable fish and wildlife populations (i.e., adequate to ensure long-term delivery of valued ecosystem services) is a widely accepted regulatory goal upon which risk assessments and risk management decisions are based.</p>
<h4>References</h4>
<ul>
<li>Miller DH, Ankley GT. 2004. Modeling impacts on populations: fathead minnow (Pimephales promelas) exposure to the endocrine disruptor 17ß-trenbolone as a case study. Ecotoxicology and Environmental Safety 59: 1-9.</li>
<h4>References</h4>
<ul>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Ankley GT, Jensen KM, Makynen EA, Kahl MD, Korte JJ, Hornung MW, Henry TR, Denny JS, Leino RL, Wilson VS, Cardon MD, Hartig PC, Gray LE. 2003. Effects of the androgenic growth promoter 17b-trenbolone on fecundity and reproductive endocrinology of the fathead minnow. Environ. Toxicol. Chem. 22: 1350–1360.</span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Bookhout TA. 1994. Research and management techniques for wildlife and habitats. The Wildlife Society, Bethesda, Maryland. 740 pp.</span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Cheng WY, Zhang Q, Schroeder A, Villeneuve DL, Ankley GT, Conolly R. 2016. Computational modeling of plasma vitellogenin alterations in response to aromatase inhibition in fathead minnows. Toxicol Sci 154: 78–89.</span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Conolly RB, Ankley GT, Cheng W-Y, Mayo ML, Miller DH, Perkins EJ, Villeneuve DL, Watanabe KH. 2017. Quantitative adverse outcome pathways and their application to predictive toxicology. Environ. Sci. Technol. 51: 4661-4672.</span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Etterson MA, Ankley GT. 2021. Endogenous Lifecycle Models for Chemical Risk Assessment. Environ. Sci. Technol. 55: 15596-15608. </span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Gotelli NJ, 2008. A Primer of Ecology. Sinauer Associates, Inc., Sunderland, MA, USA.</span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Holbech H, Kinnberg KL, Brande-Lavridsen N, Bjerregaard P, Petersen GI, Norrgren L, Orn S, Braunbeck T, Baumann L, Bomke C, Dorgerloh M, Bruns E, Ruehl-Fehlert C, Green JW, Springer TA, Gourmelon A. 2012 Comparison of zebrafish (<em>Danio rerio</em>) and fathead minnow <em>(Pimephales promelas</em>) as test species in the Fish Sexual Development Test (FSDT). Comp. Biochem. Physiol. C Toxicol. Pharmacol. 155: 407–415.</span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Kramer VJ, Etterson MA, Hecker M, Murphy CA, Roesijadi G, Spade DJ, Stromberg JA, Wang M, Ankley GT. </span><span style="color:black">2011. Adverse outcome pathways and risk assessment: Bridging to population level effects. Environ. Toxicol. Chem. 30, 64-76.</span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">McComb B, Zuckerberg B, Vesely D, Jordan C. 2021. Monitoring Animal Populations and their Habitats: A Practitioner's Guide. Pressbooks, Oregon State University, Corvallis, OR Version 1.13, 296 pp. </span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Miller DH, Villeneuve DL, Santana Rodriguez KJ, Ankley GT. 2022. A multidimensional matrix model for predicting the effect of male biased sex ratios on fish populations. Environmental Toxicology and Chemistry 41(4): 1066-1077.</span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Miller DH, Tietge JE, McMaster ME, Munkittrick KR, Xia X, Griesmer DA, Ankley GT. 2015. </span><span style="color:black">Linking mechanistic toxicology to population models in forecasting recovery from chemical stress: A case study from Jackfish Bay, Ontario, Canada. Environmental Toxicology and Chemistry 34(7): 1623-1633.</span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Miller DH, Jensen KM, Villeneuve DE, Kahl MD, Makynen EA, Durhan EJ, Ankley GT. 2007. </span><span style="color:black">Linkage of biochemical responses to population-level effects: A case study with vitellogenin in the fathead minnow (<em>Pimephales promelas</em>). Environ Toxicol Chem 26: 521–527.</span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Miller DH, Ankley GT. 2004. Modeling impacts on populations: Fathead minnow (<em>Pimephales promelas</em>) exposure to the endocrine disruptor 17b-trenbolone as a case study. Ecotox Environ Saf 59: 1–9.</span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Murphy CA, Nisbet RM, Antczak P, Garcia-Reyero N, Gergs A, Lika K, Mathews T, Muller EB, Nacci D, Peace A, Remien CH, Schultz IR, Stevenson LM, Watanabe KH. 2018. Incorporating suborganismal processes into dynamic energy budget models for ecological risk assessment. Integrated Environmental Assessment and Management 14(5): 615–624.</span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Murray DL, Sandercock BK (editors). 2020. Population ecology in practice. Wiley-Blackwell, Oxford UK, 448 pp.</span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Nisbet RM, Jusup M, Klanjscek T, Pecquerie L. 2011. Integrating dynamic energy budget (DEB) theory with traditional bioenergetic models. The Journal of Experimental Biology 215: 892-902.</span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Nisbet RM, Muller EB, Lika K, Kooijman SALM. 2000. </span><span style="color:black">From molecules to ecosystems through dynamic energy budgets. J Anim Ecol 69: 913–926.</span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Perkins EJ, Ashauer R, Burgoon L, Conolly R, Landesmann B,, Mackay C, Murphy CA, Pollesch N, Wheeler JR, Zupanic A, Scholzk S. 2019. Building and applying quantitative adverse outcome pathway models for chemical hazard and risk assessment. Environmental Toxicology and Chemistry 38(9): 1850–1865. </span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Vandermeer JH, Goldberg DE. 2003. Population ecology: first principles. Princeton University Press, Princeton NJ, 304 pp.</span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Villeneuve DL, Crump D, Garcia-Reyero N, Hecker M, Hutchinson TH, LaLone CA, Landesmann B, Lattieri T, Munn S, Nepelska M, Ottinger MA, Vergauwen L, Whelan M. Adverse outcome pathway (AOP) development 1: Strategies and principles. Toxicol Sci. 2014: 142:312–320</span></span></span></li>
<li><span style="font-size:12pt"><span style="font-family:Calibri,sans-serif"><span style="color:black">Watanabe KH, Mayo M, Jensen KM, Villeneuve DL, Ankley GT, Perkins EJ. 2016. Predicting fecundity of fathead minnows (<em>Pimephales promelas</em>) exposed to endocrine‐disrupting chemicals using a MATLAB(R)‐based model of oocyte growth dynamics. PLoS One 11: e0146594.</span></span></span></li>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt"><strong>Life stage</strong>: This KER is applicable to all life stages following the differentiation of the liver. Larvae prior to liver differentiation should not be included.</span></span></span></p>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt"><strong>Sex</strong>: This KER is applicable to both sexes.</span></span></span></p>
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<!-- if nothing shows up in any of these fields, then evidence supporting this KER will not be displayed -->
<h4>Evidence Supporting this KER</h4>
<strong>Biological Plausibility</strong>
<p><em>
High degree of plausibility in fathead minnow, zebrafish and other cyprinid species.
</em>
</p>
<strong>Empirical Evidence</strong>
<p><em>
A wide range of studies using adult fish show that induction of plasma vitellogenin (VTG) occurs within 21 days in vivo aquatic exposure to estrogen receptor agonists (eg 17beta-estradiol and 4-tert pentylphenol) as shown during the successful validation of the OECD Test Guideline 229 and related protocols. A smaller number of experiment studies with fish have shown that within the OECD Test Guideline 2010, larval fish can also show induction of whole body VTG levels within 21 days aquatic exposure to estrogen receptor agonists.
</em>
</p>
<strong>Uncertainties and Inconsistencies</strong>
<p><em>
There are generally few inconsistencies for experimental studies using model fish species dervied from pathogen-free laboratory cultures. However, there can some uncertainties where wild fish have been used for experimental purposes.
</em>
</p>
<h4>Evidence Supporting this KER</h4>
<strong>Biological Plausibility</strong>
<p>Original text - unknown contributor</p>
<p><em>High degree of plausibility in fathead minnow, zebrafish and other cyprinid species.</em></p>
<p> </p>
<p> </p>
<p><span style="color:#333399">Added by C. Baettig June 24, 2024</span></p>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">In egg laying vertebrates such as fish vitellogenin (VTG) synthesis occurs in the female liver after activation of estrogen receptors (ERs), including ERα and ERβ isoforms, by endogenous steroids and a variety of exogenous chemicals that bind to ERs (e.g., Brock & Shapiro, 1983; Denslow et al., 1999; Miracle et al., 2006). In mature female fish VTG is incorporated into growing oocytes by the ovary and is converted into yolk protein. However, neither adult male fish nor juvenile fish normally produce VTG, but the hepatic ER is present in males, as are the genes that encode for <em>vtg</em> expression and can therefore be induced by exogenous compounds (Heppell et al., 1995). </span></span></span></p>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Agonism of the ER is expected to increase <em>vtg</em> transcription and translation and under high estrogen stimulation the fold increase of <em>vtg</em> transcripts increases by orders of magnitude (Brock & Shapiro, 1983). As such, induction of VTG levels in male fish has been used extensively as a biomarker of estrogen exposure (Wheeler et al., 2005).</span></span></span></p>
<strong>Empirical Evidence</strong>
<p>Original text - unknown contributor</p>
<p><em>A wide range of studies using adult fish show that induction of plasma vitellogenin (VTG) occurs within 21 days in vivo aquatic exposure to estrogen receptor agonists (eg 17beta-estradiol and 4-tert pentylphenol) as shown during the successful validation of the OECD Test Guideline 229 and related protocols. A smaller number of experiment studies with fish have shown that within the OECD Test Guideline 2010, larval fish can also show induction of whole body VTG levels within 21 days aquatic exposure to estrogen receptor agonists. </em></p>
<p> </p>
<p> </p>
<p> </p>
<p><span style="color:#333399">Added by C. Baettig June 24, 2024</span></p>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">There are numerous publications supporting this relationship including multiple review articles (e.g., Matozzo et al., 2008; Palmer & Selcer, 1996; Verderame & Scudiero, 2017). A few specific examples are listed below.</span></span></span></p>
<ul>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Estradiol and diarylpropionitrile (DPN), an ERβ selective agonist, induced a dose-dependent increase in VTG synthesis in rainbow trout hepatocytes (Leaños-Castañeda & Van Der Kraak, 2007).</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">DPN has also been shown to increase ERα and <em>vtg</em> expression and synthesis post-injection in Mozambique tilapia in vivo (Davis et al., 2010). </span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">A study focusing on benzophenone derivatives found that BP1 (2,4-dihydroxybenzophenone), BP2 (2,2′,4,4′-tetrahydroxybenzophenone), and THB (2,4,4′-trihydroxybenzophenone) were human ERα (hERα) and hERβ and rainbow trout ERα (rtERα) and rtERβ agonists. To investigate ER activation profiles of the derivatives in vitro tests, i.e., competitive binding, reporter gene based assays, vitellogenin (Vtg) induction in isolated rainbow trout hepatocytes, and proliferation based assays were completed. hERβ was more strongly activated, which is an inverse finding to natural ligand 17β-estradiol (E2) where hERα is more strongly activated. BPs were more active in rtERα than in hERα assays. Significant VTG induction was detected in hERα, hERβ, rtERα, and rtERβ cultures (Molina-Molina et al., 2008).</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Tollefsen et al. (2003) looked at multiple endogenous (e.g., estrone (E1), estradiol (E2), and estriol (E3)) and exogenous estrogens (e.g., ethynyloestradiol (EE2), diethylstilbestrol (DES), genistein, zearalenone, bisphenol A) and found they induced dose-dependent VTG synthesis in Atlantic salmon hepatocytes.</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Shen et al. (2021) used in silico methods to screen 1056 pesticides for potential agonistic activity. They found 72 pesticides to be potential ER agonists, 14 of which have been previously reported as ER agonists. To test whether these pesticides were ER agonists, 10 were selected from the list, three that were previously reported as ER agonists and seven previously unreported as ER agonists. They found all 10 pesticides exhibited ERα agonistic activity in human or zebrafish cells and of the 10, seven also induced <em>vtg1</em> and <em>vtg2</em> mRNA in zebrafish. </span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Xu et al (2020) also showed increase in plasma VTG following exposure to aryloxy-phenoxypropionate (APP) herbicides, after measuring the binding patterns of quizalofop-P-ethyl (QPE), clodinafop-propargyl (CP) and haloxyfop-P (HP) with ERα. </span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">In male fathead minnows exposed to E2 and <span style="font-size:12.0pt">1H,1H,10H,10H-perfluorodecane-1,10-diol</span> (FC-10 diol) for 21 days expression of hepatic <em>esr1</em> and <em>vtg</em> were both significantly increased (Ankley et al. in prep).</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">In male fathead minnows exposed to methoxychlor, a weak estrogen agonist, there was a clear induction of VTG (Ankley et al. 2001). In the same study exposure to methyltestosterone, a synthetic androgen, caused a significant induction of VTG in both male and female fathead minnows. This level of induction in female fathead minnows resulted in a dose-dependent increase in VTG, to concentrations approximately 10-fold higher than those observed in control fish. These funding were likely due to the conversion of methyltestosterone to methylestradiol (Hornung et al., 2004).</span></span></span></li>
<strong>Uncertainties and Inconsistencies</strong>
<p>Original text - unknown contributor</p>
<p><em>There are generally few inconsistencies for experimental studies using model fish species dervied from pathogen-free laboratory cultures. However, there can some uncertainties where wild fish have been used for experimental purposes. </em></p>
<p> </p>
<p> </p>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif">Added by C. Baettig June 24, 2024</span></span></p>
<p> </p>
<ul>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Some uncertainty remains regarding which ER subtypes regulate <em>vtg</em> gene expression in the liver of fish. In general, the literature suggests a close interplay between ER subtypes, primarily ERα and Erβ, in the regulation of vitellogenesis. Consequently, at present, the key event relationship is generalized to impacts on all ER subtypes, even though it remains possible that impacts on a particular sub-type may drive the effect on vitellogenin transcription and translation.</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Using selective agonists and antagonists for ERα and ERβ, it was concluded that ERβ was primarily responsible for inducing vitellogenin production in rainbow trout and that compounds exhibiting ERα selectivity would not be detected using a vitellogenin ELISA bioassay (Leaños-Castañeda & Van Der Kraak, 2007). However, a subsequent study conducted in tilapia concluded that agonistic and antagonistic characteristics of mammalian, isoform-specific ER agonists and antagonists, cannot be reliably extrapolated to piscine ERs (Davis et al., 2010).</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Based on RNA interference knock-down experiments Nelson and Habibi (2010) proposed a model in which all ER subtypes are involved in E2-mediated vitellogenesis, with ERβ isoforms stimulating expression of both vitellogenin and ERα gene expression, and ERα helping to drive vitellogenesis, particularly as it becomes more abundant following sensitization.</span></span></span></li>
</ul>
<h4>References</h4>
<p><br />
Navas, J.M., Segner, H. (2006) Vitellogenin synthesis in primary cultures of fish liver cells as endpoint for in vitro screening of the (anti)estrogenic activity of chemical substances. Aquatic Toxicology 80: 1-22
</p><p>Thorpe, K.L., Benstead, R., Hutchinson, T.H., Tyler, C.R. (2007). Associations between altered vitellogenin concentrations and adverse health effects in fathead minnow (Pimephales promelas). Aquatic Toxicology 85: 176-183
</p>
<h4>References</h4>
<p>Navas, J.M., Segner, H. (2006) Vitellogenin synthesis in primary cultures of fish liver cells as endpoint for in vitro screening of the (anti)estrogenic activity of chemical substances. Aquatic Toxicology 80: 1-22</p>
<p>Thorpe, K.L., Benstead, R., Hutchinson, T.H., Tyler, C.R. (2007). Associations between altered vitellogenin concentrations and adverse health effects in fathead minnow (Pimephales promelas). Aquatic Toxicology 85: 176-183</p>
<p> </p>
<ul>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Brock, M. L., & Shapiro, D. (1983). Estrogen regulates the absolute rate of transcription of the Xenopus laevis vitellogenin genes. <em>Journal of Biological Chemistry</em>,<em> 258</em>(9), 5449-5455. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Davis, L., Katsu, Y., Iguchi, T., Lerner, D., Hirano, T., & Grau, E. (2010). Transcriptional activity and biological effects of mammalian estrogen receptor ligands on three hepatic estrogen receptors in Mozambique tilapia. <em>The Journal of steroid biochemistry and molecular biology</em>,<em> 122</em>(4), 272-278. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Denslow, N. D., Chow, M. C., Kroll, K. J., & Green, L. (1999). Vitellogenin as a biomarker of exposure for estrogen or estrogen mimics. <em>Ecotoxicology</em>,<em> 8</em>, 385-398. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Hornung, M. W., Jensen, K. M., Korte, J. J., Kahl, M. D., Durhan, E. J., Denny, J. S., Henry, T. R., & Ankley, G. T. (2004). Mechanistic basis for estrogenic effects in fathead minnow (Pimephales promelas) following exposure to the androgen 17α-methyltestosterone: conversion of 17α-methyltestosterone to 17α-methylestradiol. <em>Aquatic Toxicology</em>,<em> 66</em>(1), 15-23. <a href="https://doi.org/https:/doi.org/10.1016/j.aquatox.2003.06.004" style="color:#0563c1; text-decoration:underline">https://doi.org/https://doi.org/10.1016/j.aquatox.2003.06.004</a> </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Leaños-Castañeda, O., & Van Der Kraak, G. (2007). Functional characterization of estrogen receptor subtypes, ERα and ERβ, mediating vitellogenin production in the liver of rainbow trout. <em>Toxicology and applied pharmacology</em>,<em> 224</em>(2), 116-125. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Matozzo, V., Gagné, F., Marin, M. G., Ricciardi, F., & Blaise, C. (2008). Vitellogenin as a biomarker of exposure to estrogenic compounds in aquatic invertebrates: A review. <em>Environment International</em>,<em> 34</em>(4), 531-545. <a href="https://doi.org/https:/doi.org/10.1016/j.envint.2007.09.008" style="color:#0563c1; text-decoration:underline">https://doi.org/https://doi.org/10.1016/j.envint.2007.09.008</a> </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Miracle, A., Ankley, G., & Lattier, D. (2006). Expression of two vitellogenin genes (vg1 and vg3) in fathead minnow (Pimephales promelas) liver in response to exposure to steroidal estrogens and androgens. <em>Ecotoxicology and environmental safety</em>,<em> 63</em>(3), 337-342. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Molina-Molina, J.-M., Escande, A., Pillon, A., Gomez, E., Pakdel, F., Cavaillès, V., Olea, N., Aït-Aïssa, S., & Balaguer, P. (2008). Profiling of benzophenone derivatives using fish and human estrogen receptor-specific in vitro bioassays. <em>Toxicology and applied pharmacology</em>,<em> 232</em>(3), 384-395. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Nelson, E. R., & Habibi, H. R. (2010). Functional Significance of Nuclear Estrogen Receptor Subtypes in the Liver of Goldfish. <em>Endocrinology</em>,<em> 151</em>(4), 1668-1676. <a href="https://doi.org/10.1210/en.2009-1447" style="color:#0563c1; text-decoration:underline">https://doi.org/10.1210/en.2009-1447</a> </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Palmer, B. D., & Selcer, K. W. (1996). Vitellogenin as a biomarker for xenobiotic estrogens: a review. <em>Environmental Toxicology and Risk Assessment: Biomarkers and Risk Assessment: Fifth Volume</em>, 3-22. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Shen, C., Zhu, K., Ruan, J., Li, J., Wang, Y., Zhao, M., He, C., & Zuo, Z. (2021). Screening of potential oestrogen receptor α agonists in pesticides via in silico, in vitro and in vivo methods. <em>Environmental Pollution</em>,<em> 270</em>, 116015. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Tollefsen, K.-E., Mathisen, R., & Stenersen, J. (2003). Induction of vitellogenin synthesis in an Atlantic salmon (Salmo salar) hepatocyte culture: a sensitive in vitro bioassay for the oestrogenic and anti-oestrogenic activity of chemicals. <em>Biomarkers</em>,<em> 8</em>(5), 394-407. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Verderame, M., & Scudiero, R. (2017). Estrogen-dependent, extrahepatic synthesis of vitellogenin in male vertebrates: A mini-review. <em>Comptes Rendus Biologies</em>,<em> 340</em>(3), 139-144. <a href="https://doi.org/https:/doi.org/10.1016/j.crvi.2017.01.005" style="color:#0563c1; text-decoration:underline">https://doi.org/https://doi.org/10.1016/j.crvi.2017.01.005</a> </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Wheeler, J. R., Gimeno, S., Crane, M., Lopez-Juez, E., & Morritt, D. (2005). Vitellogenin: a review of analytical methods to detect (anti) estrogenic activity in fish. <em>Toxicology Mechanisms and Methods</em>,<em> 15</em>(4), 293-306. </span></span></li>
<li><span style="font-size:11.0pt"><span style="font-family:"Calibri",sans-serif">Xu, Y., Feng, R., Wang, L., Dong, L., Liu, R., Lu, H., & Wang, C. (2020). Computational and experimental investigations on the interactions of aryloxy-phenoxy-propionate herbicides to estrogen receptor alpha in zebrafish. <em>Ecotoxicology and environmental safety</em>,<em> 189</em>, 110003. </span></span></li>
</ul>
</div>
<br>
<div>
<div>
<h4><a href="/relationships/254">Relationship: 254: Increase, Plasma vitellogenin concentrations leads to Increase, Renal pathology due to VTG deposition</a></h4>
<td><a href="/aops/29">Estrogen receptor agonism leading to reproductive dysfunction</a></td>
<td>adjacent</td>
<td>High</td>
<td></td>
</tr>
</thead>
<tbody>
<tr>
<th><a href="/aops/29">Estrogen receptor agonism leading to reproductive dysfunction</a></th>
<th>adjacent</th>
<th>High </th>
<th></th>
</tr>
</tbody>
</table>
</div>
<h4>Evidence Supporting Applicability of this Relationship</h4>
<br>
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<div>
</div>
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<div>
</div>
<!-- end sex terms -->
<p><em>
Publish studies specifically relate to fish, although it is plausible that the same response may occur in the aquatic life-stages of amphibians.
</em>
</p>
<!-- if nothing shows up in any of these fields, then evidence supporting this KER will not be displayed -->
<h4>Evidence Supporting this KER</h4>
<strong>Biological Plausibility</strong>
<p><em>
High level of biological plausibility in fish.
</em>
</p>
<strong>Empirical Evidence</strong>
<p><em>
Laboratory in vivo aquatic exposures of fish (fathead minnow) to 17alpha-ethinylestradiol led to renal pathology within 16 weeks, concomitant with macroscopic evidence of osmoregulatory dysfunction and morbidity (Laenge et al., 2001).
</em>
</p>
<strong>Uncertainties and Inconsistencies</strong>
<p><em>
None that the author of this entry is aware of.
</em>
</p>
<tr>
<td><a href="/aops/536">Estrogen receptor agonism leading to reduced survival and population growth due to renal failure</a></td>
<td>adjacent</td>
<td>Moderate</td>
<td>Low</td>
</tr>
</tbody>
</table>
</div>
<h4>Evidence Supporting Applicability of this Relationship</h4>
<div>
</div>
<div>
</div>
<div>
</div>
<p>Original text - unknown contribution</p>
<p><em>Publish studies specifically relate to fish, although it is plausible that the same response may occur in the aquatic life-stages of amphibians.</em></p>
<p> </p>
<p><span style="color:#333399">Added by C. Baettig on June 24, 2024</span></p>
<p><span style="font-family:Arial,Helvetica,sans-serif"><span style="color:#333399"><span style="font-size:11pt"><strong>Taxonomic applicability</strong>: Oviparous vertebrates that synthesize yolk precursor proteins and have functional kidneys.</span></span></span></p>
<p><span style="font-family:Arial,Helvetica,sans-serif"><span style="color:#333399"><span style="font-size:11pt"><strong>Life stage</strong>:<strong> </strong>This KER is applicable to all life stages following the differentiation of the liver and kidney.</span></span></span></p>
<p><span style="font-family:Arial,Helvetica,sans-serif"><span style="color:#333399"><span style="font-size:11pt"><strong>Sex</strong>:<strong> </strong>This KER is applicable to both sexes.</span></span></span></p>
<h4>Evidence Supporting this KER</h4>
<strong>Biological Plausibility</strong>
<p>Original text - unknown contribution</p>
<p><em>High level of biological plausibility in fish. </em></p>
<p><span style="color:#333399">Added by C. Baettig on June 24, 2024</span></p>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">When large quantities of VTG are circulating hyalin material can accumulate in the kidneys which can cause significant pathology (Folmar et al., 2001; Herman & Kincaid, 1988; Palace et al., 2002). Additionally, eosinophilic material is known to accumulate in kidney tubules and has been proposed to be due to high circulating VTG (Hahlbeck et al., 2004). Similarly, cilia proliferation observed in renal tubules is assumed to be related to increased absorption of circulating vitellogenin (Zha et al., 2008).</span></span></span></p>
<strong>Empirical Evidence</strong>
<p>Original text - unknown contribution</p>
<p><em>Laboratory in vivo aquatic exposures of fish (fathead minnow) to 17alpha-ethinylestradiol led to renal pathology within 16 weeks, concomitant with macroscopic evidence of osmoregulatory dysfunction and morbidity (Laenge et al., 2001). </em></p>
<p><span style="color:#333399">Added by C. Baettig on June 24, 2024</span></p>
<p> </p>
<ul>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Male summer flounder injected with 17β-estradiol (E2) had increased levels of circulating VTG. The accumulation of VTG resulted in obstruction or rupture of renal glomeruli (Folmar et al., 2001).</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Male rare minnow exposed to 17α-ethinylestradiol (EE2) and 4-nonylphenol (NP) had significantly increased plasma VTG concentrations, as did females after EE2 exposure. This resulted in hemorrhages in male kidney tubules, hypertrophy of tubular epithelia, and accumulated eosinophilic material in renal tissue (Zha et al., 2007).</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Elevated levels of VTG and kidney hypertrophy in juvenile three-spined sticklebacks was observed after exposure to E2 and EE2 (Hahlbeck et al., 2004).</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Male fathead minnows experimentally exposed to EE2 within a whole lake experiment showed 9000-fold higher VTG concentrations than fish captured from the same lake prior to the EE2 additions. Edema in the interstitium between kidney tubules and eosinophilic deposits in the kidney tubule lumen were also observed in the EE2-exposed male fatheads (Palace et al., 2002).</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">After exposure to bisphenol A VTG levels increased in fathead minnows resulting in glomerular epithelial cell hyperplasia, hyaline droplets in glomeruli, glomerular mesangial membrane thickening, intravascular proteinaceous fluid, tubular dilation, and dilation of Bowman’s spaces (Mihaich et al., 2012).</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Fathead minnow embryos exposed to EE2 exhibited increased whole body VTG levels and tubular degeneration and dilation and glomerulonephritis/glomerulosclerosis was observable after 16 weeks (Länge et al., 2001).</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">In male fathead minnows exposed to E2 and an estrogenic PFAS, FC-10 diol, for 21 days plasma VTG was significantly increased. Neuropathy in the kidneys of diol-exposed fish was observed, specifically tubule dilation, tubule protein, enlarged glomeruli, glomerular protein, and thickened basement membranes. Additionally, interstitial and intravascular proteinaceous fluid was significantly elevated (Ankley et al. in prep).</span></span></span></li>
<strong>Uncertainties and Inconsistencies</strong>
<p>Original text - unknown contribution</p>
<p><em>None that the author of this entry is aware of. </em></p>
<p><span style="color:#333399">Added by C. Baettig on June 24, 2024</span></p>
<p> </p>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Although the accumulation of hyalin material/lipoprotein within the kidneys has been confirmed to be partially caused by accumulated VTG, some of the accumulated proteins do not respond to VTG antibody (e.g., Folmar et al., 2001). Because male fish will also express other estrogen inducible proteins such as vitelline envelope and zona radiata some renal pathology could be caused by these related proteins rather than VTG (Johan Hyllner et al., 1994; Oppen‐Berntsen et al., 1994).</span></span></span></p>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Proliferative kidney disease (PKD) in fish caused by the parasite <em>Tetracapsuloides bryosalmonae</em> results in significant kidney pathology. However, when PKD infection took place under simultaneous exposure to EE2, kidney pathology was less pronounced despite the fact that hepatic <em>vtg</em> was elevated in fish exposed to the estrogen (Bailey et al., 2019; Rehberger et al., 2020).</span></span></span></p>
<h4>References</h4>
<p>Herman, R.L., Kincaid, H.L. (1988) Pathological effects of orally administered 17beta-estradiol to rainbow trout. Aquaculture 72:165–172
</p><p>Länge, R., Hutchinson, T.H., Croudace, C.P., Siegmund, F., Schweinfurth, H., Hampe, P., Panter, G.H., Sumpter, J.P. (2001) Effects of the synthetic estrogen 17 alpha-ethinylestradiol on the life-cycle of the fathead minnow (Pimephales promelas). Environ Toxicol Chem 20:1216-1227
</p>
<h4>References</h4>
<p>Herman, R.L., Kincaid, H.L. (1988) Pathological effects of orally administered 17beta-estradiol to rainbow trout. Aquaculture 72:165–172</p>
<p>Länge, R., Hutchinson, T.H., Croudace, C.P., Siegmund, F., Schweinfurth, H., Hampe, P., Panter, G.H., Sumpter, J.P. (2001) Effects of the synthetic estrogen 17 alpha-ethinylestradiol on the life-cycle of the fathead minnow (Pimephales promelas). Environ Toxicol Chem 20:1216-1227</p>
<p> </p>
<ul>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Bailey, C., von Siebenthal, E. W., Rehberger, K., & Segner, H. (2019). Transcriptomic analysis of the impacts of ethinylestradiol (EE2) and its consequences for proliferative kidney disease outcome in rainbow trout (Oncorhynchus mykiss). <em>Comparative Biochemistry and Physiology Part C: Toxicology & Pharmacology</em>,<em> 222</em>, 31-48. <a href="https://doi.org/https:/doi.org/10.1016/j.cbpc.2019.04.009" style="color:#0563c1; text-decoration:underline">https://doi.org/https://doi.org/10.1016/j.cbpc.2019.04.009</a> </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Folmar, L. C., Gardner, G. R., Schreibman, M. P., Magliulo-Cepriano, L., Mills, L. J., Zaroogian, G., Gutjahr-Gobell, R., Haebler, R., Horowitz, D. B., & Denslow, N. D. (2001). Vitellogenin-induced pathology in male summer flounder (Paralichthys dentatus). <em>Aquatic Toxicology</em>,<em> 51</em>(4), 431-441. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Hahlbeck, E., Katsiadaki, I., Mayer, I., Adolfsson-Erici, M., James, J., & Bengtsson, B.-E. (2004). The juvenile three-spined stickleback (Gasterosteus aculeatus L.) as a model organism for endocrine disruption II—kidney hypertrophy, vitellogenin and spiggin induction. <em>Aquatic Toxicology</em>,<em> 70</em>(4), 311-326. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Herman, R. L., & Kincaid, H. L. (1988). Pathological effects of orally administered estradiol to rainbow trout. <em>Aquaculture</em>,<em> 72</em>(1-2), 165-172. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Johan Hyllner, S., Silvers, C., & Haux, C. (1994). Formation of the vitelline envelope precedes the active uptake of vitellogenin during oocyte development in the rainbow trout, Oncorhynchus mykiss. <em>Molecular Reproduction and Development</em>,<em> 39</em>(2), 166-175. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Länge, R., Hutchinson, T. H., Croudace, C. P., Siegmund, F., Schweinfurth, H., Hampe, P., Panter, G. H., & Sumpter, J. P. (2001). Effects of the synthetic estrogen 17α‐ethinylestradiol on the life‐cycle of the fathead minnow (Pimephales promelas). <em>Environmental Toxicology and Chemistry: An International Journal</em>,<em> 20</em>(6), 1216-1227. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Mihaich, E., Rhodes, J., Wolf, J., van der Hoeven, N., Dietrich, D., Hall, A. T., Caspers, N., Ortego, L., Staples, C., & Dimond, S. (2012). Adult fathead minnow, Pimephales promelas, partial life‐cycle reproductive and gonadal histopathology study with bisphenol A. <em>Environmental toxicology and chemistry</em>,<em> 31</em>(11), 2525-2535. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Oppen‐Berntsen, D., Olsen, S., Rong, C., Taranger, G., Swanson, P., & Walther, B. (1994). Plasma levels of eggshell zr‐proteins, estradiol‐17β, and gonadotropins during an annual reproductive cycle of Atlantic salmon (Salmo salar). <em>Journal of Experimental Zoology</em>,<em> 268</em>(1), 59-70. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Palace, V. P., Evans, R. E., Wautier, K., Baron, C., Vandenbyllardt, L., Vandersteen, W., & Kidd, K. (2002). Induction of vitellogenin and histological effects in wild fathead minnows from a lake experimentally treated with the synthetic estrogen, ethynylestradiol. <em>Water Quality Research Journal</em>,<em> 37</em>(3), 637-650. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Rehberger, K., Wernicke von Siebenthal, E., Bailey, C., Bregy, P., Fasel, M., Herzog, E. L., Neumann, S., Schmidt-Posthaus, H., & Segner, H. (2020). Long-term exposure to low 17α-ethinylestradiol (EE2) concentrations disrupts both the reproductive and the immune system of juvenile rainbow trout, Oncorhynchus mykiss. <em>Environment International</em>,<em> 142</em>, 105836. <a href="https://doi.org/https:/doi.org/10.1016/j.envint.2020.105836" style="color:#0563c1; text-decoration:underline">https://doi.org/https://doi.org/10.1016/j.envint.2020.105836</a> </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Zha, J., Sun, L., Zhou, Y., Spear, P. A., Ma, M., & Wang, Z. (2008). Assessment of 17α-ethinylestradiol effects and underlying mechanisms in a continuous, multigeneration exposure of the Chinese rare minnow (Gobiocypris rarus). <em>Toxicology and applied pharmacology</em>,<em> 226</em>(3), 298-308. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Zha, J., Wang, Z., Wang, N., & Ingersoll, C. (2007). Histological alternation and vitellogenin induction in adult rare minnow (Gobiocypris rarus) after exposure to ethynylestradiol and nonylphenol. <em>Chemosphere</em>,<em> 66</em>(3), 488-495. </span></span></li>
<td><a href="/aops/29">Estrogen receptor agonism leading to reproductive dysfunction</a></td>
<td>adjacent</td>
<td>High</td>
<td></td>
</tr>
</thead>
<tbody>
<tr>
<th><a href="/aops/29">Estrogen receptor agonism leading to reproductive dysfunction</a></th>
<th>adjacent</th>
<th>High </th>
<th></th>
</tr>
</tbody>
</table>
</div>
<!-- if nothing shows up in any of these fields, then evidence supporting this KER will not be displayed -->
<h4>Evidence Supporting this KER</h4>
<strong>Biological Plausibility</strong>
<p><em>
High level of physiological plausibility in fish.
</em>
</p>
<tr>
<td><a href="/aops/536">Estrogen receptor agonism leading to reduced survival and population growth due to renal failure</a></td>
<td>adjacent</td>
<td>High</td>
<td>Moderate</td>
</tr>
<tr>
<td><a href="/aops/537">Estrogen receptor agonism leads to reduced fecundity via increased vitellogenin in the liver</a></td>
<td>adjacent</td>
<td></td>
<td></td>
</tr>
</tbody>
</table>
</div>
<h4>Evidence Supporting Applicability of this Relationship</h4>
<div>
</div>
<div>
</div>
<div>
</div>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt"><strong>Taxonomic applicability</strong>: </span><span style="font-size:11pt">Oviparous vertebrates synthesize yolk precursor proteins that are transported in the circulation for uptake by developing oocytes. Many invertebrates also synthesize vitellogenins that are taken up into developing oocytes via active transport mechanisms. However, invertebrate vitellogenins are transported in hemolymph or via other transport mechanisms rather than plasma.</span></span></span></p>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt"><strong>Life stage</strong>: </span><span style="font-size:11pt">This KER is applicable to all life stages following the differentiation of the liver. Embryos prior to liver differentiation should not be included.</span></span></span></p>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt"><strong>Sex</strong>: </span><span style="font-size:11pt">This KER is applicable to both sexes. However, as males do not have the ability to clear plasma VTG via uptake into the oocytes the outcome is more likely to be problematic in males. Therefore, this KER has more relevance in males both in the context of monitoring for exogenous estrogens and potential biological consequences of elevated VTG.</span></span></span></p>
<h4>Evidence Supporting this KER</h4>
<strong>Biological Plausibility</strong>
<p>Original text - unknown contributor</p>
<p><em>High level of physiological plausibility in fish. </em></p>
<p> </p>
<p><span style="color:#333399">Added by C. Baettig on June 24, 2024</span></p>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">The liver is the primary source of VTG synthesis and production and after it is synthesized it is secreted into the blood (Wallace, 1985). Vitellogenin transcription and translation results in protein production, although there is a delay between expression of <em>vtg</em> and actual production/detection of VTG (e.g., Korte et al. 2000).</span></span></span></p>
<strong>Empirical Evidence</strong>
<p> </p>
<ul>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">In male tilapia, 48 hours after 17β-estradiol (E2) treatment, <em>vtg</em> hepatic mRNA expression was elevated as was plasma VTG (Davis et al., 2008).</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">In time course studies an increase in <em>vtg</em> mRNA synthesis precedes increases in plasma VTG concentration. For example, a study using male fathead minnows injected with E2, vtg mRNA was detected in the liver within 4 hours, reached a maximum around 48 hours, and returned to normal levels after 6 days. Plasma VTG was detectable within 16 hours of treatment, reached maximum levels at about 72 hours, and did not return to normal levels for at least 18 days (Korte et al., 2000).</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Similar results were observed in a flow-through experiment using sheepshead minnows exposed to E2 and p-nonylphenol. A dose dependent increase in hepatic vtg mRNA initially occurred followed by plasma VTG increase. Their results further supported that hepatic <em>vtg</em> mRNA rapidly diminishes after termination of estrogenic exposure, but plasma VTG clearance is concentration and time dependent (Hemmer et al., 2002).</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">Bowman et al. (2000) also found a time lag between <em>vtg</em>, which was elevated after 4 hours while induction of plasma VTG wasn’t detected until 24 hours in male sheepshead minnows injected with E2.</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">During waterborne exposures to 17α-ethinylestradiol (EE2), male fathead minnows showed a strong increase of <em>vtg</em> mRNA within 3 days (first sampling time point in the study), which remained elevated for the entirety of the 35-day exposure. Although plasma VTG was first detectable on day 3 it did not significantly increase until day 14 further illustrating the lag between <em>vtg</em> mRNA and plasma increase (Schmid et al., 2002).</span></span></span></li>
<li><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">In male fathead minnows exposed to E2 and FC-10 diol for 21 days, expression of hepatic <em>vtg</em> was significantly increased as was the plasma VTG (Ankley et al. in prep).</span></span></span></li>
</ul>
<strong>Uncertainties and Inconsistencies</strong>
<p><span style="font-family:Arial,Helvetica,sans-serif"><span style="color:#333399"><span style="font-size:11pt">There are no known inconsistencies between these KERs which are not readily explained on the basis of the expected dose, temporal, and incidence relationships between these two KERs. This applies across a significant body of literature in which these two KEs have been measured.</span></span></span></p>
<h4>Quantitative Understanding of the Linkage</h4>
<strong>Response-response relationship</strong>
<p><span style="font-family:Arial,Helvetica,sans-serif"><span style="color:#333399"><span style="font-size:11pt">Models and statistical relationships that define quantitative relationships between circulating E2 concentrations and circulating VTG concentrations have been developed (Ankley et al., 2008; Li et al., 2011; Murphy et al., 2009; Murphy et al., 2005). However, much of this work has focused on decreased VTG as a function of decreased E2, rather than induction.</span></span></span></p>
<strong>Time-scale</strong>
<p><span style="font-family:Arial,Helvetica,sans-serif"><span style="color:#333399"><span style="font-size:11pt">Due to the timeline between induction of mRNA transcription, translation, and the appearance of protein in plasma, as well as variable rates of uptake of VTG from plasma into oocytes, a precise quantitative relationship describing all steps of vitellogenesis transcription/translation has not been described. </span></span></span></p>
<p><span style="font-family:Arial,Helvetica,sans-serif"><span style="color:#333399"><span style="font-size:11pt">However, studies in fish suggest that that the temporal lag between mRNA transcription and increased plasma concentrations takes place within 24 hours. For example, in fish injected with E2 there is generally an increase of <em>vtg</em> mRNA beginning around 4 hours whereas plasma VTG isn’t measurable until 16-24 hours (Bowman et al., 2000; Korte et al., 2000). Additionally, in waterborne exposure of estrone (E1) in juvenile rainbow trout, elevated vtg mRNA occurred on day 4 of exposure while plasma VTG was elevated on day 5 (Osachoff et al., 2016).</span></span></span></p>
<strong>Known Feedforward/Feedback loops influencing this KER</strong>
<p><span style="color:#333399"><span style="font-family:Arial,Helvetica,sans-serif"><span style="font-size:11pt">There is no known feedback as plasma VTG does not appear to regulate expression levels in the liver.</span></span></span></p>
<h4>References</h4>
<ul>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Ankley, G. T., Miller, D. H., Jensen, K. M., Villeneuve, D. L., & Martinović, D. (2008). Relationship of plasma sex steroid concentrations in female fathead minnows to reproductive success and population status. <em>Aquatic Toxicology</em>,<em> 88</em>(1), 69-74. <a href="https://doi.org/https:/doi.org/10.1016/j.aquatox.2008.03.005" style="color:#0563c1; text-decoration:underline">https://doi.org/https://doi.org/10.1016/j.aquatox.2008.03.005</a> </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Bowman, C. J., Kroll, K. J., Hemmer, M. J., Folmar, L. C., & Denslow, N. D. (2000). Estrogen-induced vitellogenin mRNA and protein in sheepshead minnow (Cyprinodon variegatus). <em>General and Comparative Endocrinology</em>,<em> 120</em>(3), 300-313. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Davis, L. K., Pierce, A. L., Hiramatsu, N., Sullivan, C. V., Hirano, T., & Grau, E. G. (2008). Gender-specific expression of multiple estrogen receptors, growth hormone receptors, insulin-like growth factors and vitellogenins, and effects of 17β-estradiol in the male tilapia (Oreochromis mossambicus). <em>General and Comparative Endocrinology</em>,<em> 156</em>(3), 544-551. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Hemmer, M. J., Bowman, C. J., Hemmer, B. L., Friedman, S. D., Marcovich, D., Kroll, K. J., & Denslow, N. D. (2002). Vitellogenin mRNA regulation and plasma clearance in male sheepshead minnows,(Cyprinodon variegatus) after cessation of exposure to 17β-estradiol and p-nonylphenol. <em>Aquatic Toxicology</em>,<em> 58</em>(1-2), 99-112. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Korte, J. J., Kahl, M. D., Jensen, K. M., Pasha, M. S., Parks, L. G., LeBlanc, G. A., & Ankley, G. T. (2000). Fathead minnow vitellogenin: Complementary DNA sequence and messenger RNA and protein expression after 17β‐estradiol treatment. <em>Environmental Toxicology and Chemistry: An International Journal</em>,<em> 19</em>(4), 972-981. </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Li, Z., Kroll, K. J., Jensen, K. M., Villeneuve, D. L., Ankley, G. T., Brian, J. V., Sepúlveda, M. S., Orlando, E. F., Lazorchak, J. M., Kostich, M., Armstrong, B., Denslow, N. D., & Watanabe, K. H. (2011). A computational model of the hypothalamic - pituitary - gonadal axis in female fathead minnows (Pimephales promelas) exposed to 17α-ethynylestradiol and 17β-trenbolone. <em>BMC Systems Biology</em>,<em> 5</em>(1), 63. <a href="https://doi.org/10.1186/1752-0509-5-63" style="color:#0563c1; text-decoration:underline">https://doi.org/10.1186/1752-0509-5-63</a> </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Murphy, C. A., Rose, K. A., Rahman, M. S., & Thomas, P. (2009). Testing and applying a fish vitellogenesis model to evaluate laboratory and field biomarkers of endocrine disruption in Atlantic croaker (Micropogonias undulatus) exposed to hypoxia. <em>Environmental toxicology and chemistry</em>,<em> 28</em>(6), 1288-1303. <a href="https://doi.org/https:/doi.org/10.1897/08-304.1" style="color:#0563c1; text-decoration:underline">https://doi.org/https://doi.org/10.1897/08-304.1</a> </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Murphy, C. A., Rose, K. A., & Thomas, P. (2005). Modeling vitellogenesis in female fish exposed to environmental stressors: predicting the effects of endocrine disturbance due to exposure to a PCB mixture and cadmium. <em>Reproductive Toxicology</em>,<em> 19</em>(3), 395-409. <a href="https://doi.org/https:/doi.org/10.1016/j.reprotox.2004.09.006" style="color:#0563c1; text-decoration:underline">https://doi.org/https://doi.org/10.1016/j.reprotox.2004.09.006</a> </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Osachoff, H. L., Brown, L. L. Y., Tirrul, L., van Aggelen, G. C., Brinkman, F. S. L., & Kennedy, C. J. (2016). Time course of hepatic gene expression and plasma vitellogenin protein concentrations in estrone-exposed juvenile rainbow trout (Oncorhynchus mykiss). <em>Comparative Biochemistry and Physiology Part D: Genomics and Proteomics</em>,<em> 19</em>, 112-119. <a href="https://doi.org/https:/doi.org/10.1016/j.cbd.2016.02.002" style="color:#0563c1; text-decoration:underline">https://doi.org/https://doi.org/10.1016/j.cbd.2016.02.002</a> </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Schmid, T., Gonzalez-Valero, J., Rufli, H., & Dietrich, D. R. (2002). Determination of vitellogenin kinetics in male fathead minnows (Pimephales promelas). <em>Toxicology Letters</em>,<em> 131</em>(1), 65-74. <a href="https://doi.org/https:/doi.org/10.1016/S0378-4274(02)00043-7" style="color:#0563c1; text-decoration:underline">https://doi.org/https://doi.org/10.1016/S0378-4274(02)00043-7</a> </span></span></li>
<li><span style="font-size:11pt"><span style="font-family:Calibri,sans-serif">Wallace, R. A. (1985). Vitellogenesis and oocyte growth in nonmammalian vertebrates. <em>Oogenesis</em>, 127-177. </span></span></li>