Decrease, Population growth rate

13674-87-8

ASLWPAWFJZFCKF-UHFFFAOYSA-N

Tris(1,3-dichloro-2-propyl) phosphate

Tris(1,3-dichloro-2-propyl)phosphate 2-Propanol, 1,3-dichloro-, phosphate (3:1)

DTXSID9026261

PCO:0000001

PCO population of organisms

PR:000007204

PR estrogen receptor

D014819

MESH vitellogenins

PCO:0000008

PCO population growth rate

GO:0030284

GO estrogen receptor activity

GO:0010467

GO gene expression

GO:0009306

GO protein secretion

GO:0006898

GO receptor-mediated endocytosis

GO:0001555

GO oocyte growth

GO:0048599

GO oocyte development

VT:1000294

VT egg quantity

WIKI decreased

Tris(1,3-dichloropropyl)phosphate - TDCPP

2018-06-19T07:35:30

2018-06-19T07:59:12

WikiUser_22

all species

WCS_90988

common ecological species Pimephales promelas

8078

NCBI Fundulus heteroclitus

8090

NCBI Oryzias latipes

WCS_7955

common ecological species Danio rerio

WCS_90988

common ecological species fathead minnow

7955

NCBI zebra danio

8090

NCBI medaka

Decrease, Population growth rate

Population

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 assessments, population growth rate (and hence population size over time) is important to consider within the context of applied conservation practices. 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).  Equation 1:  r = b - d This key event refers to scenarios where r < 0 (instantaneous death rate exceeds instantaneous birth rate). Examining r in the context of population growth rate: ● A population will decrease to extinction when the instantaneous death rate exceeds the instantaneous birth rate (r < 0).              ● The smaller the value of r below 1, the faster the population will decrease to zero.   ● A population will increase when resources are available and the instantaneous birth rate exceeds the instantaneous death rate (r > 0)            ● The larger the value that r exceeds 1, the faster the population can increase over time       ● 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).      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.   ● 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).              Alternatively, a stressor could indirectly impact survival and/or reproduction.   ● 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) Density dependence can be an important consideration: ● 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.   ● 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).        Closed versus open systems: ● 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.   ● When individuals depart (emigrate out of a population) the loss will diminish population growth rate.   Population growth rate applies to all organisms, both sexes, and all life stages.

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, Nt=0 (population size at time t=0), and the population size at the end of the interval, Nt=1 (population size at time t = 1), and then subsequently dividing by the initial population size.  Equation 2:  r = (Nt=1 - Nt=0) / Nt=0 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).   ● 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.   ● 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).    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).   Some examples of modeling constructs used to investigate population growth rate: ● 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 (Pimephales promelas) exposed to endocrine disrupting chemicals in combination with a population model to examine projected alterations in population growth rate.   ● 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.   ● 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). ● 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. ● 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).

Consideration of population size and changes in population size over time is potentially relevant to all living organisms.

Not Specified Unspecific

Not Specified

All life stages

High

<ul> <li>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.</li> <li>Bookhout TA. 1994. Research and management techniques for wildlife and habitats. The Wildlife Society, Bethesda, Maryland. 740 pp.</li> <li>Caswell H. 2001. Matrix Population Models. Sinauer Associates, Inc., Sunderland, MA, USA</li> <li>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.</li> <li>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.</li> <li>Etterson MA, Ankley GT.  2021.  Endogenous Lifecycle Models for Chemical Risk Assessment. Environ. Sci. Technol. 55:  15596-15608. </li> <li>Gotelli NJ, 2008. A Primer of Ecology. Sinauer Associates, Inc., Sunderland, MA, USA.</li> <li>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 (Danio rerio) and fathead minnow (Pimephales promelas) as test species in the Fish Sexual Development Test (FSDT). Comp. Biochem. Physiol. C Toxicol. Pharmacol. 155:  407–415.</li> <li>Kramer VJ, Etterson MA, Hecker M, Murphy CA, Roesijadi G, Spade DJ, Stromberg JA, Wang M, Ankley GT.  2011.  Adverse outcome pathways and risk assessment: Bridging to population level effects.  Environ. Toxicol. Chem. 30, 64-76.</li> <li>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. </li> <li>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.</li> <li>Miller DH, Tietge JE, McMaster ME, Munkittrick KR, Xia X, Griesmer DA, Ankley GT. 2015. 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.</li> <li>Miller DH, Jensen KM, Villeneuve DE, Kahl MD, Makynen EA, Durhan EJ, Ankley GT. 2007. Linkage of biochemical responses to population-level effects: A case study with vitellogenin in the fathead minnow (Pimephales promelas). Environ Toxicol Chem 26:  521–527.</li> <li>Miller DH, Ankley GT. 2004. Modeling impacts on populations: Fathead minnow (Pimephales promelas) exposure to the endocrine disruptor 17b-trenbolone as a case study. Ecotox Environ Saf 59: 1–9.</li> <li>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.</li> <li>Murray DL, Sandercock BK (editors).  2020.  Population ecology in practice.  Wiley-Blackwell, Oxford UK, 448 pp.</li> <li>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.</li> <li>Nisbet RM, Muller EB, Lika K, Kooijman SALM. 2000. From molecules to ecosystems through dynamic energy budgets. J Anim Ecol 69:  913–926.</li> <li>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. </li> <li>Vandermeer JH, Goldberg DE. 2003.  Population ecology: first principles.  Princeton University Press, Princeton NJ, 304 pp.</li> <li>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</li> <li>Watanabe KH, Mayo M, Jensen KM, Villeneuve DL, Ankley GT, Perkins EJ.  2016.  Predicting fecundity of fathead minnows (Pimephales promelas) exposed to endocrine‐disrupting chemicals using a MATLAB(R)‐based model of oocyte growth dynamics. PLoS One 11:  e0146594.</li> </ul> AOPWiki 2016-11-29T18:41:24

2023-01-03T09:09:06

Antagonism, Estrogen receptor

Molecular

Site of action: The site of action for the molecular initiating event is the liver (hepatocytes). Responses at the macromolecular level: Estrogen receptor antagonists have been shown to interact with the ligand binding domain of ERs. However, those interactions occur at different contact sites than those of estrogen agonists, leading to a different conformation in the transactivation domain (Brzozowski et al. 1997; Katzenellenbogen 1996). Characterization of chemical properties: Two broad categories of ER antagonists have been described. Type I, like tamoxifen act as mixed agonists and antagonists. Type II, like ICI164384 are pure antagonists (Katzenellenbogen 1996). Due to their potential utility for treating estrogen-dependent breast cancers and other estrogen-related disease states as well as concerns regarding endocrine disruption, there is an extensive body of literature on the identification and design of chemical structures that act as ER antagonists (e.g., (Brooks et al. 1987; Brooks and Skafar 2004; Lloyd et al. 2006; Sodero et al. 2012; Vedani et al. 2012; Wang et al. 2006).

<ul> <li>The BG1luc estrogen receptor transactivation test method for identifying estrogen receptor agonists and antagonists (OECD Test Guideline 457) has been validated by the National Toxicology Program Interagency Center for Evaluation of Alternative Toxicological Methods (NICEATM) and Interagency Coordinating Committee on the Validation of Alternative Methods (ICCVAM) as an appropriate assay for detecting ER antagonism. (OECD, 2012b).</li> <li>Other human ER-based transactivation assays that have been used to detect ERα antagonism include the T47D-Kbluc assay (Wilson et al. 2004); ERα CALUX assay (van der Burg et al. 2010); MELN assay (Witters et al. 2010); and the yeast estrogen screen (YES; (De Boever et al. 2001)). Each of these assays have undergone some level of validation.</li> <li>In aquatic ecotoxicology, vitellogenin synthesis in primary fish liver cells and liver slices has also been used to screen for anti-estrogenic activity (e.g., (Bickley et al. 2009; Navas and Segner 2006; Schmieder et al. 2000; Schmieder et al. 2004; Sun et al. 2010). Although these approaches have generally not been subject to as much formal validation as human ER-based transactivation assays, in the case of fish-specific AOPs linked to this key event, these measures of anti-estrogenicity may be more directly relevant to predicting other key events in the pathway.</li> </ul>

Taxonomic applicability: Steroid receptors, including ER are thought to have evolved in the chordate lineage (Baker 1997, 2003; Thornton 2001). An ER ortholog has been isolated from a mollusk species, but no ER orthologs have been detected in arthropods or nematodes (Thornton et al. 2003). Broadly speaking, most vertebrates can be expected to have functional ERs, while most invertebrates do not, although there may be exceptions within the mollusk lineage and evolutionarily-related organisms.

CL:0000182

CL hepatocyte

<ul> <li>Brzozowski AM, Pike AC, Dauter Z, Hubbard RE, Bonn T, Engstrom O, et al. 1997. Molecular basis of agonism and antagonism in the oestrogen receptor. Nature 389:753-758.</li> <li>Katzenellenbogen B. 1996. Estrogen receptors: Bioactivities and interactions with cell signaling pathways. Biology of Reproduction 54:287-293.</li> <li>Brooks SC, Wappler NL, Corombos JD, Doherty LM, Horwitz JP. 1987. Estrogen structure-fuction relationships. Berlin:Walter de Gruyter & Co., 443-466.</li> <li>Brooks SC, Skafar DF. 2004. From ligand structure to biological activity: Modified estratrienes and their estrogenic and antiestrogenic effects in mcf-7 cells. Steroids 69:401-418.</li> <li>Lloyd DG, Smith HM, O'Sullivan T, Knox AS, Zisterer DM, Meegan MJ. 2006. Antiestrogenically active 2-benzyl-1,1-diarylbut-2-enes: Synthesis, structure-activity relationships and molecular modeling study for flexible estrogen receptor antagonists. Medicinal chemistry 2:147-168.</li> <li>Sodero AC, Romeiro NC, da Cunha EF, de Oliveira Magalhaaes U, de Alencastro RB, Rodrigues CR, et al. 2012. Application of 4d-qsar studies to a series of raloxifene analogs and design of potential selective estrogen receptor modulators. Molecules 17:7415-7439.</li> <li>Vedani A, Dobler M, Smiesko M. 2012. Virtualtoxlab - a platform for estimating the toxic potential of drugs, chemicals and natural products. Toxicology and applied pharmacology 261:142-153.</li> <li>Wang CY, Ai N, Arora S, Erenrich E, Nagarajan K, Zauhar R, et al. 2006. Identification of previously unrecognized antiestrogenic chemicals using a novel virtual screening approach. Chemical research in toxicology 19:1595-1601.</li> <li>Denny JS, Tapper MA, Schmieder PK, Hornung MW, Jensen KM, Ankley GT, et al. 2005. Comparison of relative binding affinities of endocrine active compounds to fathead minnow and rainbow trout estrogen receptors. Environmental toxicology and chemistry / SETAC 24:2948-2953.</li> <li>Lee HK, Kim TS, Kim CY, Kang IH, Kim MG, Jung KK, et al. 2012. Evaluation of in vitro screening system for estrogenicity: Comparison of stably transfected human estrogen receptor-alpha transcriptional activation (oecd tg455) assay and estrogen receptor (er) binding assay. The Journal of toxicological sciences 37:431-437.</li> <li>Rider CV, Hartig PC, Cardon MC, Lambright CR, Bobseine KL, Guillette LJ, Jr., et al. 2010. Differences in sensitivity but not selectivity of xenoestrogen binding to alligator versus human estrogen receptor alpha. Environmental toxicology and chemistry / SETAC 29:2064-2071.</li> <li>OECD. 2012b. Test no. 457: Bg1luc estrogen receptor transactivation test method for identifying estrogen receptor agonists and antagonists:OECD Publishing.</li> <li>Wilson VS, Bobseine K, Gray LE, Jr. 2004. Development and characterization of a cell line that stably expresses an estrogen-responsive luciferase reporter for the detection of estrogen receptor agonist and antagonists. Toxicological sciences : an official journal of the Society of Toxicology 81:69-77.</li> <li>van der Burg B, Winter R, Weimer M, Berckmans P, Suzuki G, Gijsbers L, et al. 2010. Optimization and prevalidation of the in vitro eralpha calux method to test estrogenic and antiestrogenic activity of compounds. Reproductive toxicology 30:73-80.</li> <li>Witters H, Freyberger A, Smits K, Vangenechten C, Lofink W, Weimer M, et al. 2010. The assessment of estrogenic or anti-estrogenic activity of chemicals by the human stably transfected estrogen sensitive meln cell line: Results of test performance and transferability. Reproductive toxicology 30:60-72.</li> <li>De Boever P, Demare 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 109:691-697.</li> <li>Bickley LK, Lange A, Winter MJ, Tyler CR. 2009. Evaluation of a carp primary hepatocyte culture system for screening chemicals for oestrogenic activity. Aquatic toxicology 94:195-203.</li> <li>Navas JM, 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.</li> <li>Schmieder P, Tapper M, Linnum A, Denny J, Kolanczyk R, Johnson R. 2000. Optimization of a precision-cut trout liver tissue slice assay as a screen for vitellogenin induction: Comparison of slice incubation techniques. Aquatic toxicology 49:251-268.</li> <li>Schmieder PK, Tapper MA, Denny JS, Kolanczyk RC, Sheedy BR, Henry TR, et al. 2004. Use of trout liver slices to enhance mechanistic interpretation of estrogen receptor binding for cost-effective prioritization of chemicals within large inventories. Environmental science & technology 38:6333-6342.</li> <li>Sun L, Wen L, Shao X, Qian H, Jin Y, Liu W, et al. 2010. Screening of chemicals with anti-estrogenic activity using in vitro and in vivo vitellogenin induction responses in zebrafish (danio rerio). Chemosphere 78:793-799.</li> <li>Baker ME. 1997. Steroid receptor phylogeny and vertebrate origins. Molecular and cellular endocrinology 135:101-107.</li> <li>Baker ME. 2003. Evolution of adrenal and sex steroid action in vertebrates: A ligand-based mechanism for complexity. BioEssays : news and reviews in molecular, cellular and developmental biology 25:396-400.</li> <li>Thornton JW. 2001. Evolution of vertebrate steroid receptors from an ancestral estrogen receptor by ligand exploitation and serial genome expansions. Proceedings of the National Academy of Sciences of the United States of America 98:5671-5676.</li> <li>Thornton JW, Need E, Crews D. 2003. Resurrecting the ancestral steroid receptor: Ancient origin of estrogen signaling. Science 301:1714-1717.</li> </ul> AOPWiki 2016-11-29T18:41:22

2017-09-16T10:14:46

Reduction, Vitellogenin synthesis in liver

Tissue

Vitellogenin is an egg yolk precursor protein synthesized by hepatocytes of oviparous vertebrates. In vertebrates, transcription of vitellogenin genes is predominantly regulated by estrogens via their action on nuclear estrogen receptors. During vitellogenic periods of the reproductive cycle, when circulating estrogen concentrations are high, vitellogenin transcription and synthesis are typically orders of magnitude greater than during non-reproductive conditions.

Relative abundance of vitellogenin transcripts or protein can be readily measured in liver tissue  (e.g., (Biales et al. 2007)) or whole body (H Holbech et al. 2001) from organisms exposed in vivo, or in liver slices (e.g., (Schmieder et al. 2000) or hepatocytes (e.g., (Navas and Segner 2006) exposed in vitro, using real-time quantitative polymerase chain reaction (PCR; transcripts) or enzyme linked immunosorbent assay (ELISA; protein).

Oviparous vertebrates. 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)

UBERON:0002107

UBERON liver

Not Specified Unspecific High Female

High

Adult, reproductively mature

High High High High

<ul> <li>Biales AD, Bencic DC, Lazorchak JL, Lattier DL. 2007. A quantitative real-time polymerase chain reaction method for the analysis of vitellogenin transcripts in model and nonmodel fish species. Environ Toxicol Chem 26(12): 2679-2686.</li> <li>Navas JM, 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. Aquat Toxicol 80(1): 1-22.</li> <li>Schmieder P, Tapper M, Linnum A, Denny J, Kolanczyk R, Johnson R. 2000. Optimization of a precision-cut trout liver tissue slice assay as a screen for vitellogenin induction: comparison of slice incubation techniques. Aquat Toxicol 49(4): 251-268.</li> <li>Wahli W. 1988. Evolution and expression of vitellogenin genes. Trends in Genetics. 4:227-232.</li> <li> H Holbech, L Andersen, G I Petersen, B Korsgaard, K L Pedersen, P Bjerregaard, Development of an ELISA for vitellogenin in whole body homogenate of zebrafish (Danio rerio), Comp Biochem Physiol C Toxicol Pharmacol. 2001, 130(1):119-31 </li> </ul> AOPWiki 2016-11-29T18:41:23

2021-05-27T01:10:51

Reduction, Plasma vitellogenin concentrations

Organ

Vitellogenin synthesized in the liver is secreted into the blood and circulates to the ovaries for uptake.

Vitellogenin concentrations in plasma are typically detected using enzyme linked Immunosorbent assay (ELISA; e.g., (Korte et al. 2000; Tyler et al. 1996; Holbech et al. 2001; Fenske et al. 2001). Although less specific and/or sensitive, determination of alkaline-labile phosphate or Western blotting has also been employed.

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.

UBERON:0001969

UBERON blood plasma

High

Adult, reproductively mature

High High High

<ul> <li>Fenske M, van Aerle R, Brack S, Tyler CR, Segner H. Development and validation of a homologous zebrafish (Danio rerio Hamilton-Buchanan) vitellogenin enzyme-linked immunosorbent assay (ELISA) and its application for studies on estrogenic chemicals. Comp Biochem Physiol C Toxicol Pharmacol. 2001. Jul;129(3):217-32.</li> <li>Holbech H, Andersen L, Petersen GI, Korsgaard B, Pedersen KL, Bjerregaard P. Development of an ELISA for vitellogenin in whole body homogenate of zebrafish (Danio rerio). Comp Biochem Physiol C Toxicol Pharmacol. 2001 Sep;130(1):119-31.</li> <li>Korte JJ, Kahl MD, Jensen KM, Mumtaz SP, Parks LG, LeBlanc GA, et al. 2000. Fathead minnow vitellogenin: complementary DNA sequence and messenger RNA and protein expression after 17B-estradiol treatment. Environmental Toxicology and Chemistry 19(4): 972-981.</li> <li>Tyler C, van der Eerden B, Jobling S, Panter G, Sumpter J. 1996. Measurement of vitellogenin, a biomarker for exposure to oestrogenic chemicals, in a wide variety of cyprinid fish. Journal of Comparative Physiology and Biology 166: 418-426.</li> <li>Wahli W. 1988. Evolution and expression of vitellogenin genes. Trends in Genetics. 4:227-232.</li> </ul> AOPWiki 2016-11-29T18:41:23

2017-09-16T10:14:37

Reduction, Vitellogenin accumulation into oocytes and oocyte growth/development

Cellular

Vitellogenin from the blood is selectively taken up by competent oocytes via receptor-mediated endocytosis. Although vitellogenin receptors mediate the uptake, opening of intercellular channels through the follicular layers to the oocyte surface as the oocyte reaches a “critical” size is thought to be a key trigger in allowing vitellogenin uptake (Tyler and Sumpter 1996). Once critical size is achieved, concentrations in the plasma and temperature are thought to impose the primary limits on uptake (Tyler and Sumpter 1996). Uptake of vitellogenin into oocytes causes considerable oocyte growth during vitellogenesis, accounting for up to 95% of the final egg size in many fish (Tyler and Sumpter 1996). Given the central role of vitellogenesis in oocyte maturation, vitellogenin accumulation is a prominent feature used in histological staging of oocytes (e.g., (Leino et al. 2005; Wolf et al. 2004).

Relative vitellogenin accumulation can be evaluated qualitatively using routine histological approaches (Leino et al. 2005; Wolf et al. 2004). Oocyte size can be evaluated qualitatively or quantitatively using routine histological and light microscopy and/or imaging approaches.

Oviparous vertebrates and invertebrates. Although hormonal regulation of vitellogenin synthesis and mechanisms of vitellogenin transport from the site of synthesis to the ovary vary between vertebrates and invertebrates (Wahli 1988), in both vertebrates and invertebrates, vitellogenin is incorporated into oocytes and cleaved to form yolk proteins.

CL:0000023

CL oocyte

High Female

High

Adult, reproductively mature

Moderate Moderate

<ul> <li>Leino R, Jensen K, Ankley G. 2005. Gonadal histology and characteristic histopathology associated with endocrine disruption in the adult fathead minnow. Environmental Toxicology and Pharmacology 19: 85-98.</li> <li>Tyler C, Sumpter J. 1996. Oocyte growth and development in teleosts. Reviews in Fish Biology and Fisheries 6: 287-318.</li> <li>Wolf JC, Dietrich DR, Friederich U, Caunter J, Brown AR. 2004. Qualitative and quantitative histomorphologic assessment of fathead minnow Pimephales promelas gonads as an endpoint for evaluating endocrine-active compounds: a pilot methodology study. Toxicol Pathol 32(5): 600-612.</li> </ul> AOPWiki 2016-11-29T18:41:23

2017-09-16T10:14:38

Reduction, Cumulative fecundity and spawning

Individual

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.

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. 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>).

Cumulative fecundity and spawning can, in theory, be evaluated for any egg laying animal.

High Female

High

Adult, reproductively mature

High High High

<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), Test No. 240: Medaka Extended One Generation Reproduction Test (MEOGRT), OECD Publishing, Paris. DOI: <a href="http://dx.doi.org/10.1787/9789264242258-en" target="_blank" title="http://dx.doi.org/10.1787/9789264242258-en">http://dx.doi.org/10.1787/9789264242258-en</a></li> <li>OECD. 2012a. Test no. 229: Fish short term reproduction assay. Paris, France:Organization for Economic Cooperation and Development.</li> </ul> AOPWiki 2016-11-29T18:41:22

2017-03-20T17:52:57

<upstream-id>b609e606-8764-457b-aaa3-a4bdc4c8220b</upstream-id> <downstream-id>c0444089-cbbd-40b4-b4ba-fefbf0769b39</downstream-id>

Vitellogenin synthesis in fish is localized in the liver and is well documented to be regulated by estrogens via interaction with estrogen receptors (Arukwe and Goksøyr 2003; Nelson and Habibi 2010; Tyler and Sumpter 1996; Tyler et al. 1996). During vitellogenic periods of the reproductive cycle, antagonism of the ER would be expected to reduce vitellogenin transcription and translation.

<ul> <li>Eleven ER antagonists were shown to reduce in vitro E2-induced vitellogenin production by rainbow trout hepatocytes in a concentration-dependent manner (Petersen and Tollefsen 2012).</li> <li>A review by Navas and Segner (Navas and Segner 2006) reports inhibition of E2-induced vitellogenin induction in primary fish liver cells to be useful for detecting anti-estrogens.</li> <li>Methyl-piperidino-pyrazole (MPP), a synthetic mammalian ERα antagonist shown to antagonize activation of the tilapia ER, also reduced the ability of estradiol to stimulate vitellogenin production in tilapia injected with the compound in vivo (Davis et al. 2010).</li> <li>A synthetic musk shown to act as a both a human and rainbow trout ERα antagonist inhibited E2-induced vitellogenin production in rainbow trout injected with the same compound (Simmons et al. 2010).</li> <li>Tamoxifen inhibited vitellogenin induction in roach (Rutilus rutilus) liver explants exposed in vitro (Gerbron et al. 2010).</li> <li>Hydroxytamoxifen inhibited vitellogenin induction in brown trout primary hepatocytes and rainbow trout and zebrafish liver cell lines (Christianson-Heiska and Isomaa 2008).</li> <li>Nafoxidine and CI-628 inhibit estrogen-induced vitellogenin synthesis in roosters (Gschwendt 1975).</li> </ul>

<ul> <li>Some uncertainty remains regarding which ER subtype(s) regulates vitellogenin gene expression in the liver of fish. In general, the literature suggests a close interplay between several ER subtypes 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.</li> </ul> <ul> <li>Griffin et al. reported that morpholino knock-downs of either esr1 (ERα) or esr2b (ERβb) prevented estradiol-mediated induction of vitellogenin expression in zebrafish (Griffin et al. 2013).</li> <li>Using selective agonists 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 bioassay (Leanos-Castaneda and 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).</li> <li>Based on RNA interference knock-down experiments Nelson and Habibi 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 (Nelson and Habibi 2010).</li> </ul>

At present, we are not aware of any studies that have defined the quantitative relationship between measures of ER antagonism and the magnitude or severity of impaired vitellogenin transcription or translation in liver.

#<Reference::ActiveRecord_Associations_CollectionProxy:0x00007b42ea24d1e8> AOPWiki 2016-11-29T18:41:33

2016-11-30T12:10:50

<upstream-id>c0444089-cbbd-40b4-b4ba-fefbf0769b39</upstream-id> <downstream-id>a2bf9a2a-79da-4aca-baf6-d7b44dce62e1</downstream-id> See biological plausibility, below.

Updated 03/20/2017.

Liver is the major source of VTG protein production in fish (Tyler and Sumpter 1996; Arukwe and Goksøyr 2003). Protein production involves transcription and subsequent translation. The time-lag between decreases in transcription/translation and decreases in plasma VTG concentrations can be expected to be dependent on vitellogenin elimination half-lives.

<ul> <li>In a number of time-course experiments with aromatase inhibitors, decreases in plasma estradiol concentrations precede decreases in plasma vitellogenin concentrations (Villeneuve et al. 2009; Skolness et al. 2011; Ankley et al. 2009b). Recovery of plasma E2 concentrations also precedes recovery of plasma VTG concentrations after cessation of exposure (Villeneuve et al. 2009; Ankley et al. 2009a; Villeneuve et al. 2013).</li> <li>In experiments with strong estrogens, increases in vtg mRNA synthesis precede increases in plasma VTG concentration (Korte et al. 2000; Schmid et al. 2002).</li> <li>Elimination half-lives for VTG protein have been determined for induced male fish, but to our knowledge, similar kinetic studies have not been done for reproductively mature females (Korte et al. 2000; Schultz et al. 2001).</li> <li>In male sheepshead minnows injected with E2, induction of VTG mRNA precedes induction of plasma VTG (Bowman et al. 2000).</li> <li>In male Cichlasoma dimerus exposed to octylphenol for 28 days and then held in clean water, decline in induced VTG mRNA concentrations precedes declines in induced plasma VTG concentrations (Genovese et al. 2012).</li> </ul>

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.

Due to temporal disconnects (lag) between induction of mRNA transcription and translation and significant changes in plasma concentrations as well as variable rates of uptake of VTG from plasma into oocytes, a precise quantitative relationship between VTG transcription/translation and circulating VTG concentrations has not been described. However, models and statistical relationships that define quantitative relationships between circulating E2 concentrations and circulating VTG concentrations have been developed (Li et al. 2011a; Murphy et al. 2005; Murphy et al. 2009; Ankley et al. 2008).

Not Specified Unspecific

Not Specified

Adult, reproductively mature

Moderate

This KER primarily applies to taxa that synthesize vitellogenin in the liver which is transported elsewhere in the body via plasma (i.e., oviparous vertebrates).

#<Reference::ActiveRecord_Associations_CollectionProxy:0x00007b42ea29b4b0> AOPWiki 2016-11-29T18:41:33

2017-03-20T12:58:16

<upstream-id>a2bf9a2a-79da-4aca-baf6-d7b44dce62e1</upstream-id> <downstream-id>ba0dc139-faf7-4d2c-b207-d276874cf1a3</downstream-id> SEE BIOLOGICAL PLAUSIBILITY BELOW

Vitellogenin synthesized in the liver and transported to the ovary via the circulation is the primary source of egg yolk proteins in fish (Wallace and Selman 1981; Tyler and Sumpter 1996; Arukwe and Goksøyr 2003). In many teleosts vitellogenesis can account for up to 95% of total egg size (Tyler and Sumpter 1996).

In some (Ankley et al. 2002; Ankley et al. 2003; Lalone et al. 2013), but not all (Ankley et al. 2005; Sun et al. 2007; Skolness et al. 2013) fish reproduction studies, reductions in plasma vitellogenin have been associated with visible decreases in yolk protein content in oocytes and overall reductions in ovarian stage.

Not all fish reproduction studies showing reductions in plasma vitellogenin have caused visible decreases in yolk protein content in oocytes and overall reductions in ovarian stage. (Ankley et al. 2005; Sun et al. 2007; Skolness et al. 2013). While plasma vitellogenin is well established as the only major source of vitellogenins to the oocyte, the extent to which a decrease will impact an ovary that has already developed vitellogenic staged oocytes is less certain. It would be assumed that the more rapid the turn-over of oocytes in the ovary, the tighter the linkage between these KEs. Thus, repeat spawning species with asynchronous oocyte development that spawn frequently would likely be more vulnerable than annual spawning species with synchronous oocyte development that had already reached late vitellogenic stages.

<ul> <li>Rates of vitellogenin uptake as a function of ovarian follicle surface area have been estimated for rainbow trout, an annual spawning fish species, and may exceed 700 ng/mm2 follicle surface per hour (Tyler and Sumpter 1996).</li> <li>Comparable data are lacking for repeat-spawning species and kinetic relationships between plasma concentrations and uptake rates within the ovary have not been defined.</li> <li>A model based on a statistical relationship between plasma E2 concentrations, spawning interval, and cumulative fecundity has been developed to predict changes in cumulative fecundity from plasma VTG (Li et al. 2011b), but it does not incorporate a model of the kinetics of VTG uptake nor the influence of VTG uptake on oocyte growth.</li> </ul>

High Female

High

Adult, reproductively mature

Moderate Moderate

This KER is expected to be primarily applicable to oviparous vertebrates that synthesize vitellogenin in hepatic tissue which is ultimately incorporated into oocytes present in the ovary.

#<Reference::ActiveRecord_Associations_CollectionProxy:0x00007b42ea42f970> AOPWiki 2016-11-29T18:41:33

2017-03-20T13:21:09

<upstream-id>ba0dc139-faf7-4d2c-b207-d276874cf1a3</upstream-id> <downstream-id>c0197bfc-bed4-48b6-9bdb-f1674cb896f2</downstream-id> SEE BIOLOGICAL PLAUSIBILITY BELOW

Vitellogenesis is a critical stage of oocyte development and accumulated lipids and yolk proteins make up the majority of oocyte biomass (Tyler and Sumpter 1996). At least in mammals, maintenance of meiotic arrest is supported by signals transmitted through gap junctions between the granulosa cells and oocytes (Jamnongjit and Hammes 2005). Disruption of oocyte-granulosa contacts as a result of cell growth has been shown to coincide with oocyte maturation (Eppig 1994). However, it remains unclear whether the relationship between vitellogenin accumulation and oocyte growth and eventual maturation is causal or simply correlative.

<ul> <li>At present, to our best knowledge there are no studies that definitively demonstrate a direct cause-effect relationship between impaired VTG accumulation into oocytes and impaired spawning. There is, however, strong correlative evidence. Across a range of laboratory studies with small fish, there is a robust and statistically significant correlation between reductions in circulating VTG concentrations and reductions in cumulative fecundity (Miller et al. 2007). To date, we are unaware of any fish reproduction studies which show a large reduction in circulating VTG concentrations, but not reductions in cumulative fecundity.</li> <li>Ankley et al. (2003) reported significant reductions in VTG accumulation in oocytes along with significant reductions in cumulative fecundity, although fecundity was significantly impacted at a lower dose (0.05 ug/L 17beta-trenbolone versus 0.5 ug/L for VTG accumulation).</li> <li>Kang et al. (2008) reported significant reductions in both VTG accumulation in occytes and cumulative fecundity in Japanese medaka, with cumulative fecundity being impacted at slightly lower concentrations (0.047 ug 17alpha-methyltestosterone/L versus 0.088 ug/L).</li> <li> </li> </ul>

Based on the limited number of studies available that have examined both of these KEs, there are no known, unexplained, results that are inconsistent with this relationship.

Across a range of laboratory studies with fathead minnow, there is a robust and statistically significant correlation between reductions in circulating VTG concentrations and reductions in cumulative fecundity (Miller et al. 2007). At present it is unclear how well that relationship may hold for other fish species or feral fish under the influence of environmental variables. A model based on a statistical relationship between plasma E2 concentrations, spawning interval, and cumulative fecundity has been developed to predict changes in cumulative fecundity from plasma VTG (Li et al. 2011b). However, to date, such models do not specifically consider vitellogenin uptake into oocytes as a quantitative predictor of fecundity. Furthermore, with the exception of a few specialized studies, quantitative measures of VTG content in oocytes are rarely measured in toxicity studies. In contrast, plasma VTG is routinely measured.

High Female

High

Adult, reproductively mature

Moderate Moderate

On the basis of the taxonomic relevance of the two KEs linked via this KER, this KER is likely applicable to aquatic, oviparous, vertebrates which both produce vitellogenin and deposit eggs/sperm into an aquatic environment.

#<Reference::ActiveRecord_Associations_CollectionProxy:0x00007b42ea50b1f0> AOPWiki 2016-11-29T18:41:33

2017-03-20T13:35:29

<upstream-id>c0197bfc-bed4-48b6-9bdb-f1674cb896f2</upstream-id> <downstream-id>fd763f98-584a-472f-acfb-3184e69b8e6b</downstream-id> SEE BIOLOGICAL PLAUSIBILITY BELOW

Updated 03/20/2017

Using a relatively simple density-dependent population model and assuming constant young of year survival with no immigration/emigration, reductions in cumulative fecundity have been predicted to yield declines in population size over time (Miller and Ankley 2004). Under real-world environmental conditions, outcomes may vary depending on how well conditions conform with model assumptions. Nonetheless, cumulative fecundity can be considered one vital rate that contributes to overall population trajectories (Kramer et al. 2011).

<ul> <li>Using a relatively simple density-dependent population model and assuming constant young of year survival with no immigration/emigration, reductions in cumulative fecundity have been predicted to yield declines in population size over time (Miller and Ankley 2004). However, it should be noted that the model was constructed in such a way that predicted population size is dependent on cumulative fecundity, therefore this is a fairly weak form of empirical support.</li> <li>In a study in which an entire lake was treated with 17alpha-ethynyl estradiol, Kidd et al. (2007) declines in fathead minnow population size were associated with signs of reduced fecundity.</li> </ul>

<ul> <li>Wester et al. (2003) and references cited therein suggest that although egg production is an endpoint of demographic significance, incomplete reductions of egg production may not translate in a simple manner to population reductions. Compensatory effects of reduced predation and reduced competition for limited food and/or habitat resources may offset the effects of incomplete reductions in egg production.</li> <li>Fish and other egg laying animals employ a diverse range of reproductive strategies and life histories. The nature of the relationship between reduced spawning frequency and cumulative fecundity and overall population trajectories will depend heavily on the life history and reproductive strategy of the species in question. Relationships developed for one species will not necessarily hold for other species, particularly those with differing life histories.</li> </ul>

<ul> <li>Cumulative fecundity is one example of a vital rate that can influence population size over time. A variety of population model constructs can be adapted to utilize measurements or estimates of cumulative fecundity as a predictor of population trends over time (e.g., (Miller and Ankley 2004; Miller et al. 2013).</li> <li>The model of Miller et al. 20014 uses a relatively simple density-dependent population model and assuming constant young of year survival with no immigration/emigration, use measures of cumulative fecundity to predict relative change in in population size over time (Miller and Ankley 2004).</li> </ul>

Not Specified Unspecific

Not Specified

All life stages

Moderate

Spawning generally refers to the release of eggs and/or sperm into water, generally by aquatic or semi-aquatic organisms. Consequently, by definition, this KER is likely applicable only to organisms that spend a portion of their life-cycle in or near aquatic environments.

#<Reference::ActiveRecord_Associations_CollectionProxy:0x00007b42ea5e7c68> AOPWiki 2016-11-29T18:41:33

2017-03-20T13:49:05

Estrogen receptor antagonism leading to reproductive dysfunction

Evgeniia Kazymova

Daniel L. Villeneuve, US EPA Mid-Continent Ecology Division (villeneuve.dan@epa.gov)

BY-SA

EAGMST Under Review

1.12

1.0

This adverse outcome pathway details the linkage between antagonism of estrogen receptor in females and the adverse effect of reduced cumulative fecundity in repeat-spawning fish species. 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 (OECD 2012a). Cumulative fecundity is one of several variables known to be of demographic significance in forecasting fish population trends. Therefore, this AOP has utility in supporting the application of measures of ER antagonism, or in silico predictions of the ability to antagonize ER as a means to identify chemicals with known potential to adversely affect fish populations.

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.

adjacent

Low

High adjacent

Moderate

High adjacent

Low

Moderate adjacent

Moderate

Moderate adjacent

Moderate

High Female

High

Adult, reproductively mature

Not Specified Not Specified Not Specified

<ul> <li>Overall Assessment of the AOP</li> </ul> <ul> <li>Concordance of dose-response relationships:</li> </ul> <ul> <li>In a 42 d static renewal exposure to tamoxifen, significant, concentration dependent reduction in the number of clutches and cumulative fecundity were observed for zebrafish (Wester et al. 2003).</li> <li>A concentration-dependent reduction in circulating vitellogenin concentrations was detected in female medaka exposed to tamoxifen for 21 d (Sun et al. 2007b). Vitellogenin reductions occurred at a lower concentration (i.e., ≥ 25 μg tamoxifen/L) than reductions in fecundity (i.e., 625 μg tamoxifen/L).</li> </ul> <ul> <li>Temporal concordance among the key events and adverse effect: To date, there are no time-course studies that allow for robust evaluation of the temporal concordance of the entire AOP. However, the temporal concordance of some of the key event relationships has been established. Specifically, reductions in transcription of vitellogenin mRNAs have been shown to precede changes in circulating vitellogenin concentrations.</li> <li>Consistency:</li> </ul> <ul> <li>In zebrafish exposed to tamoxifen, reductions in the number of clutches and cumulative egg production were predicted to result in population reductions, although this was in conjunction with altered sex ratios as a concurrent effect in a partial life-cycle test (Wester et al. 2003).</li> <li>In medaka co-exposed to 17β-estradiol (E2; 200 ng/L) and 10, 50, or 250 μg tamoxifen/L, exposure to 250 μg tamoxifen significantly reduced fecundity compared to both controls and fish exposed to E2 alone (Sun et al. 2009).</li> <li>Fecundity was significantly reduced in medaka exposed to 625 μg tamoxifen/L (Sun et al. 2007b).</li> <li>Increases in atretic oocytes and oviducts filled with degenerated eggs were observed in female zebrafish exposed to tamoxifen (Wester et al. 2003). Reduced vitellogenin immuno staining was observed in tamoxifen-exposed zebrafish, based on blind semi-quantitative scoring (van der Ven et al. 2007; Wester et al. 2003). The results are therefore consistent with the AOP.</li> <li>In Japanese medaka co-exposed to E2 and tamoxifen for 21 d, both plasma vitellogenin and fecundity were reduced in a tamoxifen concentration-dependent manner (Sun et al. 2009). Although from a co-exposure, the results are broadly consistent with the AOP.</li> <li>In Japanese medaka exposed to tamoxifen for 21 d, plasma vitellogenin in females was reduced in a concentration-dependent manner and cumulative fecundity was reduced at the maximum concentration tested (Sun et al. 2007b). The results are consistent with the AOP.</li> <li>Dietary exposure to tamoxifen was also shown to reduce circulating vitellogenin concentrations in female medaka (Chikae et al. 2004). The results are consistent with the AOP.</li> <li>In tilapia co-injected with E2 or o,p-DDT, tamoxifen inhibited the stimulatory effects of E2 and o,p-DDT on plasma vitellogenin (measured as alkaline labile phosphorous). Alkaline labile phosphorous was not reduced following injection with tamoxifen alone (Leanos-Castaneda et al. 2002). These results are neither entirely consistent nor inconsistent with the AOP.</li> </ul> <ul> <li>Uncertainties, inconsistencies, and data gaps:</li> </ul> <ul> <li>In a 42 d in vivo, flow through, exposures to tamoxifen citrate, no significant reductions in circulating vitellogenin or cumulative fecundity were detected (Williams et al. 2007). The results are therefore inconsistent with the AOP.</li> <li>Some uncertainty remains regarding which ER subtype(s) regulates vitellogenin gene expression in the liver of fish. In general, the literature suggests a close interplay between several ER subtypes in the regulation of vitellogenesis. Consequently, at present, the AOP is generalized to impacts on all ER subtypes, even though it remains possible that impacts on a particular sub-type may drive the adverse response.</li> </ul> <ul> <li>Griffin et al. reported that morpholino knock-downs of either esr1 (ERα) or esr2b (ERβb) prevented estradiol-mediated induction of vitellogenin expression in zebrafish (Griffin et al. 2013).</li> <li>Using selective agonists 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 bioassay (Leanos-Castaneda and 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).</li> <li>Expression of both ERα1 and ERβ1 were strongly correlated with plasma vitellogenin concentrations over the reproductive cycle of rainbow trout (Nagler et al. 2012).</li> <li>Based on RNA interference knock-down experiments Nelson and Habibi 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 (Nelson and Habibi 2010).</li> </ul> <ul> <li>There remains uncertainty as to whether there is a direct biological linkage, as opposed to correlation only, between impaired VTG uptake into oocytes and impaired spawning/reduced cumulative fecundity. Plausible biological connections have been hypothesized but have not yet been tested experimentally.</li> </ul>   Life Stage: This AOP applies to sexually mature animals. Sex: This AOP applies to females. Taxonomic Applicability: Based on the taxonomic applicability of the component key events, this AOP could potentially apply to most oviparous chordates. Domain(s) of Applicability <ul> <li>Sex: The AOP applies to females only</li> <li>Life stages: The relevant life stages for this AOP are reproductively mature adults. This AOP does not apply to adult stages that lack a sexually mature ovary, for example as a result of seasonal or environmentally-induced gonadal senescence (i.e., through control of temperature, photo-period, etc. in a laboratory setting).</li> <li>Taxonomic: At present, the assumed taxonomic applicability domain of this AOP is class Osteichthyes. In all likelihood, the AOP will also prove applicable to all classes of fish (e.g., Agnatha and Chondrithyes as well). Additionally, all the key events described should be conserved among all oviparous vertebrates, suggesting that the AOP may also have relevance for amphibians, reptiles, and birds. However, species-specific differences in reproductive strategies/life histories, ADME (adsorption, distribution, metabolism, and elimination), compensatory reproductive endocrine responses may influence the outcomes, particularly from a quantitative standpoint.</li> </ul>

The weight of evidence for each of the KERs comprising the AOP are ranked moderate to strong. Biological plausibility at the molecular and cellular level of the early key events is very strong. Some uncertainties regarding the mechanistic details of the connection between reduced vtg availability and uptake limit the strength of evidence to some degree. However, there are considerable evidence to support the idea that ER antagonism can ultimately lead to reproductive failure. Overall weight of evidence is moderate.

A quantitative relationship between ER antagonism (the MIE) and reductions in vitellogenin transcription and translation have not been well established. However, a correlative relationship between plasma vitellogenin concentrations and cumulative fecundity has been reported (Miller et al. 2007) and applied for quantitative modeling (Ankley et al.

<ul> <li>OECD. 2012a. Test no. 229: Fish short term reproduction assay. Paris, France:Organization for Economic Cooperation and Development.</li> <li>Wester P, van den Brandhof E, Vos J, van der Ven L. 2003. Identification of endocrine disruptive effects in the aquatic environment - a partial life cycle assay in zebrafish. (RIVM Report). Bilthoven, the Netherlands:Joint Dutch Environment Ministry</li> <li>Sun L, Zha J, Spear PA, Wang Z. 2007b. Tamoxifen effects on the early life stages and reproduction of japanese medaka (oryzias latipes). Environmental toxicology and pharmacology 24:23-29.</li> <li>Sun L, Zha J, Wang Z. 2009. Effects of binary mixtures of estrogen and antiestrogens on japanese medaka (oryzias latipes). Aquatic toxicology 93:83-89.</li> <li>Williams TD, Caunter JE, Lillicrap AD, Hutchinson TH, Gillings EG, Duffell S. 2007. Evaluation of the reproductive effects of tamoxifen citrate in partial and full life-cycle studies using fathead minnows (pimephales promelas). Environmental toxicology and chemistry / SETAC 26:695-707.</li> <li>van der Ven LT, van den Brandhof EJ, Vos JH, Wester PW. 2007. Effects of the estrogen agonist 17beta-estradiol and antagonist tamoxifen in a partial life-cycle assay with zebrafish (danio rerio). Environmental toxicology and chemistry / SETAC 26:92-99.</li> <li>Chikae M, Ikeda R, Hasan Q, Morita Y, Tamiya E. 2004. Effects of tamoxifen, 17alpha-ethynylestradiol, flutamide, and methyltestosterone on plasma vitellogenin levels of male and female japanese medaka (oryzias latipes). Environmental toxicology and pharmacology 17:29-33.</li> <li>Leanos-Castaneda O, Van Der Kraak G. 2007. Functional characterization of estrogen receptor subtypes, eralpha and erbeta, mediating vitellogenin production in the liver of rainbow trout. Toxicology and applied pharmacology 224:116-125.</li> <li>Griffin LB, January KE, Ho KW, Cotter KA, Callard GV. 2013. Morpholino mediated knockdown of eralpha, erbetaa and erbetab mrnas in zebrafish (danio rerio) embryos reveals differential regulation of estrogen-inducible genes. Endocrinology.</li> <li>Davis LK, Katsu Y, Iguchi T, Lerner DT, Hirano T, Grau EG. 2010. Transcriptional activity and biological effects of mammalian estrogen receptor ligands on three hepatic estrogen receptors in mozambique tilapia. The Journal of steroid biochemistry and molecular biology 122:272-278.</li> <li>Nagler JJ, Cavileer TD, Verducci JS, Schultz IR, Hook SE, Hayton WL. 2012. Estrogen receptor mrna expression patterns in the liver and ovary of female rainbow trout over a complete reproductive cycle. General and comparative endocrinology 178:556-561.</li> <li>Nelson ER, Habibi HR. 2010. Functional significance of nuclear estrogen receptor subtypes in the liver of goldfish. Endocrinology 151:1668-1676.</li> <li>Miller DH, Jensen KM, Villeneuve DL, Kahl MD, Makynen EA, Durhan EJ, Ankley GT. 2007. Linkage of biochemical responses to population-level effects: a case study with vitellogenin in the fathead minnow (Pimephales promelas). Environ. Toxicol. Chem. 26: 521-527.</li> </ul> AOPWiki 2016-11-29T18:41:16

2023-09-25T16:26:45