102676-47-1CLPFFLWZZBQMAO-UHFFFAOYNA-NCLPFFLWZZBQMAO-UHFFFAOYSA-N
FadrozoleDTXSID5034141112809-51-5HPJKCIUCZWXJDR-UHFFFAOYSA-NHPJKCIUCZWXJDR-UHFFFAOYSA-N
LetrozoleFemera
Benzonitrile, 4,4'-(1H-1,2,4-triazol-1-ylmethylene)bis-
DTXSID402320267747-09-5TVLSRXXIMLFWEO-UHFFFAOYSA-NTVLSRXXIMLFWEO-UHFFFAOYSA-N
Prochloraz1H-Imidazole-1-carboxamide, N-propyl-N-[2-(2,4,6-trichlorophenoxy)ethyl]-
BTS 40542-7877
N-propil-N-[2-(2,4,6-triclorofenoxi)etil]-1H-imidazol-1-carboxamida
N-propyl-N-[2-(2,4,6-trichlorophenoxy)ethyl]-1H-imidazole-1-carboxamide
N-Propyl-N-[2-(2,4,6-trichlorophenoxy)ethyl-1H-imidazole-1-carboxamide
N-Propyl-N-[2-(2,4,6-trichlorphenoxy)ethyl]-1H-imidazol-1-carboxamid
Plocloraz
Prelude
Sportak
Sportake
DTXSID402427013674-87-8ASLWPAWFJZFCKF-UHFFFAOYSA-NASLWPAWFJZFCKF-UHFFFAOYSA-N
Tris(1,3-dichloro-2-propyl) phosphateTris(1,3-dichloro-2-propyl)phosphate
2-Propanol, 1,3-dichloro-, phosphate (3:1)
DTXSID9026261PCO:0000001population of organismsPR:000006100aromataseCHEBI:1646917beta-estradiolD014819vitellogeninsPCO:0000008population growth rateGO:0070330aromatase activityGO:0010467gene expressionGO:0009306protein secretionGO:0006898receptor-mediated endocytosisGO:0001555oocyte growthGO:0048599oocyte developmentGO:0006703estrogen biosynthetic processVT:1000294egg quantity2decreasedFadrozole2016-11-29T18:42:172016-11-29T18:42:17Letrozole2016-11-29T18:42:172016-11-29T18:42:17Prochloraz2016-11-29T18:42:222016-11-29T18:42:22Tris(1,3-dichloropropyl)phosphate - TDCPP2018-06-19T07:35:302018-06-19T07:59:12WikiUser_22all speciesWikiUser_28Vertebrates10116ratWCS_9606humanWCS_90988fathead minnow8078Fundulus heteroclitusWCS_90988Pimephales promelas8090Oryzias latipesWCS_7955Danio rerio8090medakaWCS_7955zebrafish10090mouseDecrease, Population growth rateDecrease, Population growth ratePopulation<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>
<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="margin-left:96px; text-align:start"><span style="font-size:medium"><span style="font-family:Calibri,sans-serif"><span style="color:#000000"><span style="color:black">Equation 2: r = (N</span><sub><span style="font-size:9pt"><span style="color:black">t=1 </span></span></sub><span style="color:black">- N</span><sub><span style="font-size:9pt"><span style="color:black">t=0</span></span></sub><span style="color:black">) / N</span><sub><span style="font-size:9pt"><span style="color:black">t=0</span></span></sub></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>
<p>Consideration of population size and changes in population size over time is potentially relevant to all living organisms.</p>
Not SpecifiedUnspecificNot SpecifiedAll life stagesHigh<ul>
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</ul>
2016-11-29T18:41:242023-01-03T09:09:06Inhibition, AromataseInhibition, AromataseMolecular<p>Inhibition of cytochrome P450 aromatase (CYP19; specifically cyp19a1a in fish).</p>
<p>Site of action: The site of action for the molecular initiating event is the ovarian granulosa cells.</p>
<p>While many vertebrates have a single isoform of aromatase, fish are known to have two isoforms. CYP19a1a is predominantly expressed in ovary while cyp19a1b is predominantly expressed in brain (Callard et al. 2001; Cheshenko et al. 2008). For the purposes of this MIE, when applied to fish, the assumed effect is on cyp19a1a. However, given that both isoforms show similar sensitivity to aromatase inhibitors (Hinfray et al. 2006) and catalyze the same reaction, discrimination of specific isoforms is not viewed as critical in relative to determining downstream key events resulting from aromatase inhibition in ovarian granulosa cells.</p>
<p>Responses at the macromolecular level: Aromatase catalyzes three sequential oxidation steps (i.e., KEGG reactions R02501, R04761, R03087 or R01840, R04759, R02351; <a class="external free" href="http://www.genome.jp/kegg/pathway.html" rel="nofollow" target="_blank">http://www.genome.jp/kegg/pathway.html</a>) involved in the conversion of C-19 androgens (e.g., testosterone, androstenedione) to C-18 estrogens (e.g., 17β-estradiol, estrone). Aromatase inhibitors interfere with one or more of these reactions, leading to reduced efficiency in converting C-19 androgens into C-18 estrogens. Therefore, inhibition of aromatase activity results in decreased rate of 17β-estradiol (and presumably estrone) production by the ovary.</p>
<p>Measurement/detection: Aromatase activity is typically measured by evaluating the production of tritiated water released upon the aromatase catalyzed conversion of radio-labeled androstenedione to estrone (Lephart and Simpson 1991). Aromatase activity can be measured in cell lines exposed in vitro (e.g., human placental JEG-3 cells and JAR choriocarcinoma cells, (Letcher et al. 1999); H295R human adrenocortical carcinoma cells (Sanderson et al. 2000)). Aromatase activity can also be quantified in tissue (i.e., ovary or brain) from vertebrates exposed in vivo (e.g., (Villeneuve et al. 2006; Ankley et al. 2002). In vitro aromatase assays are amenable to high throughput and have been included in nascent high throughput screening programs like the US EPA ToxcastTM program. Specific ToxCast assays indicative of potential aromatase inhibition include:</p>
<h1><a href="https://comptox.epa.gov/dashboard/assay-endpoints/NVS_ADME_hCYP19A1"><span style="font-size:12px">NVS_ADME_hCYP19A1</span></a></h1>
<h1><a href="https://comptox.epa.gov/dashboard/assay-endpoints/ERF_ENZ_hCYP19A1_dn"><span style="font-size:12px">ERF_ENZ_hCYP19A1_dn</span></a></h1>
<h1><a href="https://comptox.epa.gov/dashboard/assay-endpoints/TOX21_Aromatase_Inhibition"><span style="font-size:12px">TOX21_Aromatase_Inhibition</span></a></h1>
<p><strong>Taxonomic applicability</strong>: Aromatase (CYP19) orthologs are known to be present among most of the vertebrate lineage, at least down to the cartilaginous fishes. Orthologs have generally not been found in invertebrates, however, CYP19 was detected in the invertebrate chordate, amphioxus and analysis of conservation of gene order and content suggests a possible origin among primitive chordates (Castro et al. 2005).</p>
<p>Fishes generally have two aromatase isoforms, cyp19a1a which is predominantly expressed in ovary and cyp19b, predominantly expressed in brain (Callard et al. 2001). Given that cyp19a1a is dominant isoform expressed in ovary and both isoforms appear to show similar sensitivity to aromatase inhibitors (Hinfray et al., 2006), for the purpose of this key event which focuses on gonadal aromatase activty, distinction of effects on one isoform versus the other are considered negligible. Total activity, without regard to isoform can be considered.</p>
<p>Life stage applicability: Aromatase activity can be measured at any life stage after the onset of endogenous steroid biosynthesis, generally shortly after birth or hatch.</p>
<p>Sex applicability: Although expression and activity tends to be greater in females, aromatase activity can be measured in both male and female vertebrates. </p>
UBERON:0012186ovary growing follicleCL:0000501granulosa cellNot SpecifiedUnspecificNot SpecifiedAll life stagesModerate<ul>
<li>Petkov PI, Temelkov S, Villeneuve DL, Ankley GT, Mekenyan OG. 2009. Mechanism-based categorization of aromatase inhibitors: a potential discovery and screening tool. SAR QSAR Environ Res 20(7-8): 657-678.</li>
</ul>
<ul>
<li>Lephart ED, Simpson ER. 1991. Assay of aromatase activity. Methods Enzymol 206: 477-483.</li>
</ul>
<ul>
<li>Letcher RJ, van Holsteijn I, Drenth H-J, Norstrom RJ, Bergman A, Safe S, et al. 1999. Cytotoxicity and aromatase (CYP19) activity modulation by organochlorines in human placental JEG-3 and JAR choriocarcinoma cells. Toxico App Pharm 160: 10-20.</li>
</ul>
<ul>
<li>Sanderson J, Seinen W, Giesy J, van den Berg M. 2000. 2-chloro-triazine herbicides induce aromatase (CYP19) activity in H295R human adrenocortical carcinoma cells: a novel mechanism for estrogenicity. Toxicol Sci 54: 121-127.</li>
</ul>
<ul>
<li>Villeneuve DL, Knoebl I, Kahl MD, Jensen KM, Hammermeister DE, Greene KJ, et al. 2006. Relationship between brain and ovary aromatase activity and isoform-specific aromatase mRNA expression in the fathead minnow (Pimephales promelas). Aquat Toxicol 76(3-4): 353-368.</li>
</ul>
<ul>
<li>Ankley GT, Kahl MD, Jensen KM, Hornung MW, Korte JJ, Makynen EA, et al. 2002. Evaluation of the aromatase inhibitor fadrozole in a short-term reproduction assay with the fathead minnow (Pimephales promelas). Toxicol Sci 67: 121-130.</li>
</ul>
<ul>
<li>Castro LF, Santos MM, Reis-Henriques MA. 2005. The genomic environment around the Aromatase gene: evolutionary insights. BMC Evol Biol 5: 43.</li>
</ul>
<ul>
<li>Callard GV, Tchoudakova AV, Kishida M, Wood E. 2001. Differential tissue distribution, developmental programming, estrogen regulation and promoter characteristics of cyp19 genes in teleost fish. J Ster Biochem Mol Biol 79: 305-314.</li>
</ul>
<ul>
<li>Cheshenko K, Pakdel F, Segner H, Kah O, Eggen RI. Interference of endocrine disrupting chemicals with aromatase CYP19 expression or activity, and consequences for reproduction of teleost fish. Gen Comp Endocrinol. 2008 Jan 1;155(1):31-62.</li>
</ul>
<ul>
<li>Hinfray N, Porcher JM, Brion F. Inhibition of rainbow trout (Oncorhynchus mykiss) P450 aromatase activities in brain and ovarian microsomes by various environmental substances. Comp Biochem Physiol C Toxicol Pharmacol. 2006 Nov;144(3):252-62</li>
</ul>
2016-11-29T18:41:222022-03-14T08:43:09Reduction, Plasma 17beta-estradiol concentrationsReduction, Plasma 17beta-estradiol concentrationsOrgan<p>Estradiol synthesized by the gonads is transported to other tissues via blood circulation. The gonads are generally considered to be the primary source of estrogens in systemic circulation.</p>
<p>Total concentrations of 17β-estradiol in plasma can be measured by radioimmunoassay (e.g., (Jensen et al. 2001)), enzyme-linked immunosorbent assay (available through many commercial vendors), or by analytical chemistry (e.g., LC/MS; Owen et al. 2014). Total steroid hormones are typically extracted from plasma or serum via liquid-liquid or solid phase extraction prior to analysis.</p>
<p>Given that there are numerous genes, like those coding for vertebrate vitellogenins, choriongenins, cyp19a1b, etc. which are known to be regulated by estrogen response elements, targeted qPCR or proteomic analysis of appropriate targets could also be used as an indirect measure of reduced circulating estrogen concentrations. However, further support for the specificity of the individual gene targets for estrogen-dependent regulation should be established in order to support their use.</p>
<p>A line of transgenic zebrafish employing green fluorescence protein under control of estrogen response elements could also be used to provide direct evidence of altered estrogen, with decreased GFP signal in estrogen responsive tissues like liver, ovary, pituitary, and brain indicating a reduction in circulating estrogens (Gorelick and Halpern 2011).</p>
<p>Key enzymes needed to synthesize 17β-estradiol first appear in the common ancestor of amphioxus and vertebrates (Baker 2011). Consequently, this key event is applicable to most vertebrates.</p>
UBERON:0001969blood plasmaNot SpecifiedUnspecificHighAdultHighHighHighHigh<ul>
<li>Jensen K, Korte J, Kahl M, Pasha M, Ankley G. 2001. Aspects of basic reproductive biology and endocrinology in the fathead minnow (Pimephales promelas). Comparative Biochemistry and Physiology Part C 128: 127-141.</li>
<li>Baker ME. 2011. Origin and diversification of steroids: co-evolution of enzymes and nuclear receptors. Molecular and cellular endocrinology 334(1-2): 14-20.</li>
<li>Owen LJ, Wu FC, Keevil BG. 2014. A rapid direct assay for the routine measurement of oestradiol and oestrone by liquid chromatography tandem mass spectrometry. Ann. Clin. Biochem. 51(pt 3):360-367.</li>
<li>Gorelick DA, Halpern ME. Visualization of estrogen receptor transcriptional activation in zebrafish. Endocrinology. 2011 Jul;152(7):2690-703. doi: 10.1210/en.2010-1257. Epub 2011 May 3. PubMed PMID: 21540282</li>
</ul>
2016-11-29T18:41:232017-09-26T11:30:57Reduction, Vitellogenin synthesis in liverReduction, Vitellogenin synthesis in liverTissue<p>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.</p>
<p>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).</p>
<p>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)</p>
UBERON:0002107liverNot SpecifiedUnspecificHighFemaleHighAdult, reproductively matureHighHighHighHigh<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>
<p><span style="font-size:16px"><span style="color:#212121"><span style="font-family:"Segoe UI",sans-serif">H Holbech, L Andersen, </span></span><span style="color:#212121"><span style="font-family:"Segoe UI",sans-serif">G I Petersen</span></span><span style="color:#212121"><span style="font-family:"Segoe UI",sans-serif">, </span></span><span style="color:#212121"><span style="font-family:"Segoe UI",sans-serif">B Korsgaard</span></span><span style="color:#212121"><span style="font-family:"Segoe UI",sans-serif">, </span></span><span style="color:#212121"><span style="font-family:"Segoe UI",sans-serif">K L Pedersen, P Bjerregaard, </span><span style="font-family:맑은 고딕">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</span></span></span></p>
</li>
</ul>
2016-11-29T18:41:232021-05-27T01:10:51Reduction, Vitellogenin accumulation into oocytes and oocyte growth/developmentReduction, Vitellogenin accumulation into oocytes and oocyte growth/developmentCellular<p>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).</p>
<p>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.</p>
<p>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.</p>
CL:0000023oocyteHighFemaleHighAdult, reproductively matureModerateModerate<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>
2016-11-29T18:41:232017-09-16T10:14:38Reduction, 17beta-estradiol synthesis by ovarian granulosa cellsReduction, 17beta-estradiol synthesis by ovarian granulosa cellsCellular<p>Like all steroids, estradiol is a cholesterol derivative. Estradiol synthesis in ovary is mediated by a number of enzyme catalyzed reactions involving cyp11 (cholesterol side chain cleavage enzyme), cyp 17 (17alpha-hydroxylase/17,20-lyase), 3beta hydroxysteroid dehyrogenase, 17beta hydroxysteroid dehydrogenase, and cyp19 (aromatase). Among those enzyme catalyzed reactions, conversion of testosterone to estradiol, catalyzed by aromatase, is considered to be rate limiting for estradiol synthesis. Within the ovary, aromatase expression and activity is primarily localized in the granulosa cells (reviewed in (Norris 2007; Yaron 1995; Havelock et al. 2004) and others). Reactions involved in synthesis of C-19 androgens are primarily localized in the theca cells and C-19 androgens diffuse from the theca into granulosa cells where aromatase can catalyze their conversion to C-18 estrogens.</p>
<p>Due to the importance of both theca and granulosa cells in ovarian steroidogenesis, it is generally impractical to measure E2 production by isolated granulosa cells (Havelock et al. 2004). However, this key event can be evaluated by examining E2 production by intact ovarian tissue explants either exposed to chemicals in vitro (e.g., (Villeneuve et al. 2007; McMaster ME 1995) or in vivo (i.e., via ex vivo steroidogenesis assay; e.g., (Ankley et al. 2007)). Estradiol released by ovarian tissue explants into media can be quantified by radioimmunoassay (e.g., Jensen et al. 2001), ELISA, or analytical methods such as LC-MS (e.g., Owen et al. 2014).</p>
<p>OECD TG 456 <a class="external text" href="http://www.oecd-ilibrary.org/environment/test-no-456-h295r-steroidogenesis-assay_9789264122642-en" rel="nofollow" target="_blank">(OECD 2011)</a> is the validated test guideline for an in vitro screen for chemical effects on steroidogenesis, specifically the production of 17ß-estradiol (E2) and testosterone (T).</p>
<p>The synthesis of E2 can be measured in vitro cultured ovarian cells. The methods for culturing mammalian ovarian cells can be found in the Database Service on Alternative Methods to animal experimentation (DB-ALM): Culture of Human Cumulus Granulosa Cells <a class="external text" href="http://ecvam-dbalm.jrc.ec.europa.eu/beta/index.cfm/methodsAndProtocols/index?id_prot=266" rel="nofollow" target="_blank">(EURL ECVAM Protocol No. 92)</a>, Granulosa and Theca Cell Culture Systems <a class="external text" href="http://ecvam-dbalm.jrc.ec.europa.eu/beta/index.cfm/methodsAndProtocols/index?id_met=535" rel="nofollow" target="_blank">(EURL ECVAM Method Summary No. 92)</a>.</p>
<p>Key enzymes needed to synthesize 17β-estradiol first appear in the common ancestor of amphioxus and vertebrates (Markov et al. 2009; Baker 2011). Consequently, it is plausible that this key event is applicable to most vertebrates. This key event is not applicable to invertebrates, which lack the enzymes required to synthesize 17ß-estradiol.</p>
<p> </p>
CL:0000501granulosa cellHighFemaleHighAdult, reproductively matureHighHigh<ul>
<li>Ankley GT, Jensen KM, Kahl MD, Makynen EA, Blake LS, Greene KJ, et al. 2007. Ketoconazole in the fathead minnow (Pimephales promelas): reproductive toxicity and biological compensation. Environ Toxicol Chem 26(6): 1214-1223.</li>
<li>Baker ME. 2011. Origin and diversification of steroids: co-evolution of enzymes and nuclear receptors. Molecular and cellular endocrinology 334(1-2): 14-20.</li>
<li>EURL ECVAM Method Summary no 92. Granulosa and Theca Cell Culture Systems - Summary</li>
<li>EURL ECVAM Protocol no 92 Culture of Human Cumulus Granulosa Cells. Primary cell culture method. Contact Person: Dr. Mahadevan Maha M.</li>
<li>Havelock JC, Rainey WE, Carr BR. 2004. Ovarian granulosa cell lines. Molecular and cellular endocrinology 228(1-2): 67-78.</li>
<li>Jensen K, Korte J, Kahl M, Pasha M, Ankley G. 2001. Aspects of basic reproductive biology and endocrinology in the fathead minnow (Pimephales promelas). Comparative Biochemistry and Physiology Part C 128: 127-141.</li>
<li>McMaster ME MK, Jardine JJ, Robinson RD, Van Der Kraak GJ. 1995. Protocol for measuring in vitro steroid production by fish gonadal tissue. Canadian Technical Report of Fisheries and Aquatic Sciences 1961 1961: 1-78.</li>
<li>Norris DO. 2007. Vertebrate Endocrinology. Fourth ed. New York: Academic Press.</li>
<li>OECD (2011), Test No. 456: H295R Steroidogenesis Assay, OECD Guidelines for the Testing of Chemicals, Section 4, OECD Publishing, Paris.DOI: <a href="http://dx.doi.org/10.1787/9789264122642-en" target="_blank">http://dx.doi.org/10.1787/9789264122642-en</a></li>
<li>Owen LJ, Wu FC, Keevil BG. 2014. A rapid direct assay for the routine measurement of oestradiol and oestrone by liquid chromatography tandem mass spectrometry. Ann. Clin. Biochem. 51(pt 3):360-367.</li>
<li>Villeneuve DL, Ankley GT, Makynen EA, Blake LS, Greene KJ, Higley EB, et al. 2007. Comparison of fathead minnow ovary explant and H295R cell-based steroidogenesis assays for identifying endocrine-active chemicals. Ecotoxicol Environ Saf 68(1): 20-32.</li>
<li>Villeneuve DL, Mueller ND, Martinovic D, Makynen EA, Kahl MD, Jensen KM, et al. 2009. Direct effects, compensation, and recovery in female fathead minnows exposed to a model aromatase inhibitor. Environ Health Perspect 117(4): 624-631.</li>
<li>Yaron Z. 1995. Endocrine control of gametogenesis and spawning induction in the carp. Aquaculture 129: 49-73.</li>
</ul>
2016-11-29T18:41:222017-09-16T10:14:21Reduction, Cumulative fecundity and spawningReduction, Cumulative fecundity and spawningIndividual<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>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>
<p>Cumulative fecundity and spawning can, in theory, be evaluated for any egg laying animal.</p>
HighFemaleHighAdult, reproductively matureHighHighHigh<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 />
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>
2016-11-29T18:41:222017-03-20T17:52:57Reduction, Plasma vitellogenin concentrationsReduction, Plasma vitellogenin concentrationsOrgan<p>Vitellogenin synthesized in the liver is secreted into the blood and circulates to the ovaries for uptake.</p>
<p>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.</p>
<p>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.</p>
UBERON:0001969blood plasmaHighAdult, reproductively matureHighHighHigh<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>
2016-11-29T18:41:232017-09-16T10:14:3755cca9a6-3e01-4373-aba0-339db0b80bba7c0a87d2-653b-46b7-8fa5-a3e9fb9fcb5d<p>Within the ovary, aromatase expression and activity is primarily localized in the granulosa cells (reviewed in (Norris 2007; Yaron 1995; Havelock et al. 2004) and others). C-19 androgens diffuse from the theca cells into granulosa cells where aromatase can catalyze their conversion to C-18 estrogens. Therefore, inhibition of ovarian aromatase activity can generally be assumed to directly impact E2 synthesis by the granulosa cells.</p>
<h2> </h2>
<p> </p>
<ul>
<li>Known aromatase inhibitors including fadrozole and prochloraz were shown to cause concentration-dependent inhibition of aromatase activity in fathead minnow ovary homogenates (Villeneuve et al. 2006; Ankley et al. 2005).</li>
<li>Fadrozole and prochloraz also cause concentration-dependent decreases in E2 production by fathead minnow ovary explants exposed in vitro (Villeneuve et al. 2007).</li>
<li>Following in vivo exposure to fadrozole or prochloraz, ex vivo E2 production is significantly decreased in a concentration-dependent manner early in the time-course following exposure, although depending on the concentration, compensatory responses may offset the direct impact later in the exposure time-course (Villeneuve et al. 2006; Villeneuve et al. 2009; Ankley et al. 2009a; Skolness et al. 2011).</li>
</ul>
<p>Based on the limited set of studies available to date, there are no known inconsistencies.</p>
<p>Several mechanistically-based models of ovarian steroidogenesis have been developed (Breen et al. 2013; Breen et al. 2007; Shoemaker et al. 2010; Quignot and Bois 2013).</p>
<ul>
<li>The Breen et al. 2007 model was developed based on in vitro experiments with fathead minnow ovary tissue, and considers effects on steroidogenesis within the ovary only.</li>
<li>The Breen et al. 2013 model was developed based on in vivo time-course data for fathead minnow and incorporates prediction of compensatory responses resulting from feedback mechanisms operating as part of the hypothalamic-pituitary-gonadal axis.</li>
<li>The Shoemaker et al. 2010 model is chimeric and includes signaling pathways and aspects of transcriptional regulation based on a mixture of fish-specific and mammalian sources.</li>
<li>The Quignot and Bois 2013 model was designed to predict rat ovarian steroid secretion based on in vitro experiments with endocrine disrupting chemicals.</li>
</ul>
<p>These may be adaptable to predict in vitro E2 production and/or plasma E2 concentrations from in vitro or in vivo measurements of aromatase inhibition.</p>
<p>Aromatase (CYP19) orthologs are known to be present among most of the vertebrate lineage, at least down to the cartilaginous fishes. Orthologs have generally not been found in invertebrates, however, CYP19 was detected in the invertebrate chordate, amphioxus and analysis of conservation of gene order and content suggests a possible origin among primitive chordates (Castro et al. 2005).</p>
<ul>
<li>Norris DO. 2007. Vertebrate Endocrinology. Fourth ed. New York: Academic Press.</li>
<li>Yaron Z. 1995. Endocrine control of gametogenesis and spawning induction in the carp. Aquaculture 129: 49-73.</li>
<li>Havelock JC, Rainey WE, Carr BR. 2004. Ovarian granulosa cell lines. Molecular and cellular endocrinology 228(1-2): 67-78.</li>
<li>Villeneuve DL, Knoebl I, Kahl MD, Jensen KM, Hammermeister DE, Greene KJ, et al. 2006. Relationship between brain and ovary aromatase activity and isoform-specific aromatase mRNA expression in the fathead minnow (Pimephales promelas). Aquat Toxicol 76(3-4): 353-368.</li>
<li>Ankley GT, Jensen KM, Durhan EJ, Makynen EA, Butterworth BC, Kahl MD, et al. 2005. Effects of two fungicides with multiple modes of action on reproductive endocrine function in the fathead minnow (Pimephales promelas). Toxicol Sci 86(2): 300-308.</li>
<li>Villeneuve DL, Ankley GT, Makynen EA, Blake LS, Greene KJ, Higley EB, et al. 2007. Comparison of fathead minnow ovary explant and H295R cell-based steroidogenesis assays for identifying endocrine-active chemicals. Ecotoxicol Environ Saf 68(1): 20-32.</li>
<li>Villeneuve DL, Mueller ND, Martinovic D, Makynen EA, Kahl MD, Jensen KM, et al. 2009. Direct effects, compensation, and recovery in female fathead minnows exposed to a model aromatase inhibitor. Environ Health Perspect 117(4): 624-631.</li>
<li>Ankley GT, Bencic D, Cavallin JE, Jensen KM, Kahl MD, Makynen EA, et al. 2009a. Dynamic nature of alterations in the endocrine system of fathead minnows exposed to the fungicide prochloraz. Toxicol Sci 112(2): 344-353.</li>
<li>Skolness SY, Durhan EJ, Garcia-Reyero N, Jensen KM, Kahl MD, Makynen EA, et al. 2011. Effects of a short-term exposure to the fungicide prochloraz on endocrine function and gene expression in female fathead minnows (Pimephales promelas). Aquat Toxicol 103(3-4): 170-178.</li>
<li>Breen M, Villeneuve DL, Ankley GT, Bencic DC, Breen MS, Watanabe KH, et al. 2013. Developing Predictive Approaches to Characterize Adaptive Responses of the Reproductive Endocrine Axis to Aromatase Inhibition: II. Computational Modeling. Toxicological sciences : an official journal of the Society of Toxicology.</li>
<li>Breen MS, Villeneuve DL, Breen M, Ankley GT, Conolly RB. 2007. Mechanistic computational model of ovarian steroidogenesis to predict biochemical responses to endocrine active compounds. Annals of biomedical engineering 35(6): 970-981.</li>
<li>Shoemaker JE, Gayen K, Garcia-Reyero N, Perkins EJ, Villeneuve DL, Liu L, et al. 2010. Fathead minnow steroidogenesis: in silico analyses reveals tradeoffs between nominal target efficacy and robustness to cross-talk. BMC systems biology 4: 89.</li>
<li>Quignot N, Bois FY. 2013. A computational model to predict rat ovarian steroid secretion from in vitro experiments with endocrine disruptors. PloS one 8(1): e53891.</li>
</ul>
2016-11-29T18:41:332016-11-30T13:27:217c0a87d2-653b-46b7-8fa5-a3e9fb9fcb5dbee27495-701a-44ac-8330-4372a7b5c15f<p>See plausibility, below.</p>
<p>Updated 03/20/2017.</p>
<p>While brain, interrenal, adipose, and breast tissue (in mammals) are capable of synthesizing estradiol, the gonads are generally considered the major source of circulating estrogens in vertebrates, including fish (Norris 2007). Consequently, if estradiol synthesis by ovarian granulosa cells is reduced, plasma E2 concentrations would be expected to decrease unless there are concurrent reductions in the rate of E2 catabolism. Synthesis in other tissues generally plays a paracrine role only, thus the contribution of other tissues to plasma E2 concentrations can generally be considered negligible.</p>
<h2> </h2>
<p><em>Include consideration of temporal concordance here </em></p>
<p><strong>Fish</strong></p>
<ul>
<li>In multiple studies with aromatase inhibitors (e.g., fadrozole, prochloraz), significant reductions in ex vivo E2 production have been linked to, and shown to precede, reductions in circulating E2 concentrations (Villeneuve et al. 2009; Skolness et al. 2011). It is also notable that compensatory responses at the level of ex vivo steroid production (i.e., rate of E2 synthesis per unit mass of tissue) tend to precede recovery of plasma E2 concentrations following an initial insult (Villeneuve et al. 2009; Ankley et al. 2009a; Villeneuve et al. 2013).</li>
<li>Ex vivo E2 production by ovary tissue collected from female fish exposed to 30 or 300 μg ketoconazole/L showed significant decreases prior to significant effects on plasma estradiol being observed (Ankley et al. 2012).</li>
<li>Ekman et al. (2011) reported significant reductions in ex vivo E2 production and plasma E2 concentrations in female fathead minnows exposed to 0.05 ug/L 17ß-trenbolone. The effect on plasma E2 was observed at an earlier time point (24 h, versus 48 h for E2 production.</li>
<li>Rutherford et al. (2015) reported significant reductions in both E2 production and circulating E2 concentrations in female Fundulus heteroclitus exposed to 5alpha-dihydrotestosterone or 17alpha-methyltestosterone for 14 d. The effects were equipotent in the case of 17alpha-methyltestosterone, but in the case of 5alpha-dihyrotestosterone, the effect on plasma E2 could be detected at a lower dose (10 ug/L) than that at which a significant effect on E2 production was detected (100 ug/L).</li>
<li>In female Fundulus heteroclitus exposed to 17alpha-methyltestosterone for 7 or 14 d, both E2 production and plasma E2 were impacted at the same exposure concentrations (Sharpe et al. 2004). </li>
</ul>
<p><strong>Mammals</strong></p>
<ul>
<li>MEHP /DEHP, mice, ex vivo DEHP (10 -100 μg/ml); MEHP (0.1 and 10 μg/ml) dose dependent reduction E2 production (Gupta et al., 2010)</li>
<li>DEHP, rat, in vivo 300-600 mg/kg/day, dose dependent reduction of E2 plasma levels (Xu et al., 2010)</li>
</ul>
<p>Evidence for rodent and human models is summarized in Table 1.</p>
<p> </p>
<table border="1" style="border-collapse:collapse; font-size:75%">
<tbody>
<tr>
<td>
<p>Compound class</p>
</td>
<td>
<p>Species</p>
</td>
<td>
<p>Study type</p>
</td>
<td>
<p>Dose</p>
</td>
<td>
<p>E2 production/levels</p>
</td>
<td>
<p>Reference</p>
</td>
</tr>
<tr>
<td>
<p>Phthalates (DEHP)</p>
</td>
<td>
<p>rat</p>
</td>
<td>
<p>ex vivo</p>
</td>
<td>
<p>1500 mg/kg/day</p>
</td>
<td>
<p>Reduced/increased E2 production in ovary culture</p>
</td>
<td>
<p>(Laskey & Berman, 1993)</p>
</td>
</tr>
<tr>
<td>
<p>Phthalates (MEHP)</p>
</td>
<td>
<p>rat</p>
</td>
<td>
<p>in vitro</p>
</td>
<td>
<p>From 50 µM</p>
</td>
<td>
<p>Reduced E2 production (concentration and time dependent in Granulosa cell)</p>
</td>
<td>
<p>(Davis, Weaver, Gaines, & Heindel, 1994)</p>
</td>
</tr>
<tr>
<td>
<p>Phthalates (MEHP)</p>
</td>
<td>
<p>rat</p>
</td>
<td>
<p>in vitro</p>
</td>
<td>
<p>100-200µM</p>
</td>
<td>
<p>reduction E2 production (dose dependent)</p>
</td>
<td>
<p>(Lovekamp & Davis, 2001)</p>
</td>
</tr>
<tr>
<td>
<p>Phthalates (DEHP)</p>
</td>
<td>
<p>rat</p>
</td>
<td>
<p>in vivo</p>
</td>
<td>
<p>300-600 mg/kg/day</p>
</td>
<td>
<p>reduction E2 levels dose dependent</p>
</td>
<td>
<p>(Xu et al., 2010),</p>
</td>
</tr>
<tr>
<td>
<p>Phthalates (MEHP)</p>
</td>
<td>
<p>human</p>
</td>
<td>
<p>in vitro</p>
</td>
<td>
<p>IC(50)= 49- 138 µM (dependent on the stimulant)</p>
</td>
<td>
<p>reduction E2 production (dose dependent)</p>
</td>
<td>
<p>(Reinsberg, Wegener-Toper, van der Ven, van der Ven, & Klingmueller, 2009)</p>
</td>
</tr>
<tr>
<td>
<p>Phthalates (MEHP/DEHP)</p>
</td>
<td>
<p>mice</p>
</td>
<td>
<p>ex vivo</p>
</td>
<td>
<p>DEHP (10 -100 μg/ml); MEHP (0.1 and 10 μg/ml)</p>
</td>
<td>
<p>reduction E2 production (dose dependent)</p>
</td>
<td>
<p>(Gupta et al., 2010)</p>
</td>
</tr>
</tbody>
</table>
<p><br />
Table 1. Summary of the experimental data for decrease E2 production and decreased E2 levels. IC50- half maximal inhibitory concentration values reported if available, otherwise the concentration at which the effect was observed.</p>
<p>Based on the limited set of studies available to date, there are no known inconsistencies.</p>
<p>At present we are unaware of any well established quantitative relationships between ex vivo E2 production (as an indirect measure of granulosa cell E2 synthesis) and plasma E2 concentrations.</p>
<p>There are considerable data available which might support the development of such a relationship. Additionally, there are a number of existing mathematical/computational models of ovarian steroidogenesis (Breen et al. 2013; Shoemaker et al. 2010) and/or physiologically-based pharmacokinetic models of the hypothalamic-pituitary-gonadal axis (e.g., (Li et al. 2011a) that may be adaptable to support a quantitative understanding of this linkage.</p>
<p>• The Breen et al. 2013 model was developed based on in vivo time-course data for fathead minnow and incorporates prediction of compensatory responses resulting from feedback mechanisms operating as part of the hypothalamic-pituitary-gonadal axis.</p>
<p>• The Shoemaker et al. 2010 model is chimeric and includes signaling pathways and aspects of transcriptional regulation based on a mixture of fish-specific and mammalian sources.</p>
<p>• The Li et al. 2011 model is a PBPK-based model that was calibrated from data from fathead minnows, including controls and fish exposed to either 17alpha ethynylestradiol or 17beta trenbolone.</p>
HighFemaleHighAdult, reproductively matureNot SpecifiedModerateHighModerateHigh<p>Key enzymes needed to synthesize 17β-estradiol first appear in the common ancestor of amphioxus and vertebrates (Baker 2011). While some E2 synthesis can occur in other tissues, the ovary is recognized as the major source of 17β-estradiol synthesis in female vertebrates. Endocrine actions of ovarian E2 are facilitated through transport via the plasma. Consequently, this key event relationship is applicable to most female vertebrates.</p>
<ul>
<li>Ankley GT, Bencic DC, Cavallin JE, Jensen KM, Kahl MD, Makynen EA, et al. 2009a. Dynamic nature of alterations in the endocrine system of fathead minnows exposed to the fungicide prochloraz. Toxicological sciences : an official journal of the Society of Toxicology 112(2): 344-353.</li>
<li>Ankley GT, Cavallin JE, Durhan EJ, Jensen KM, Kahl MD, Makynen EA, et al. 2012. A time-course analysis of effects of the steroidogenesis inhibitor ketoconazole on components of the hypothalamic-pituitary-gonadal axis of fathead minnows. Aquatic toxicology 114-115: 88-95.</li>
<li>Baker ME. 2011. Origin and diversification of steroids: co-evolution of enzymes and nuclear receptors. Molecular and cellular endocrinology 334(1-2): 14-20.</li>
</ul>
<ul>
<li>Davis, B J, R Weaver, L J Gaines, and J J Heindel. 1994. “Mono-(2-Ethylhexyl) Phthalate Suppresses Estradiol Production Independent of FSH-cAMP Stimulation in Rat Granulosa Cells.” Toxicology and Applied Pharmacology 128 (2) (October): 224–8. doi:10.1006/taap.1994.1201.</li>
</ul>
<ul>
<li>Ekman DR, Villeneuve DL, Teng Q, Ralston-Hooper KJ, Martinović-Weigelt D, Kahl MD, Jensen KM, Durhan EJ, Makynen EA, Ankley GT, Collette TW. Use of gene expression, biochemical and metabolite profiles to enhance exposure and effects assessment of the model androgen 17β-trenbolone in fish. Environ Toxicol Chem. 2011 Feb;30(2):319-29. doi: 10.1002/etc.406.</li>
</ul>
<ul>
<li>Gupta, Rupesh K, Jeffery M Singh, Tracie C Leslie, Sharon Meachum, Jodi a Flaws, and Humphrey H-C Yao. 2010. “Di-(2-Ethylhexyl) Phthalate and Mono-(2-Ethylhexyl) Phthalate Inhibit Growth and Reduce Estradiol Levels of Antral Follicles in Vitro.” Toxicology and Applied Pharmacology 242 (2) (January 15): 224–30. doi:10.1016/j.taap.2009.10.011.</li>
</ul>
<ul>
<li>Laskey, J.W., and E. Berman. 1993. “Steroidogenic Assessment Using Ovary Culture in Cycling Rats: Effects of Bis (2-Diethylhexyl) Phthalate on Ovarian Steroid Production.” Reproductive Toxicology 7 (1) (January): 25–33. doi:10.1016/0890-6238(93)90006-S.</li>
</ul>
<ul>
<li>Li Z, Kroll KJ, Jensen KM, Villeneuve DL, Ankley GT, Brian JV, et al. 2011a. A computational model of the hypothalamic: pituitary: gonadal axis in female fathead minnows (Pimephales promelas) exposed to 17alpha-ethynylestradiol and 17beta-trenbolone. BMC systems biology 5: 63.</li>
</ul>
<ul>
<li>Lovekamp, T N, and B J Davis. 2001. “Mono-(2-Ethylhexyl) Phthalate Suppresses Aromatase Transcript Levels and Estradiol Production in Cultured Rat Granulosa Cells.” Toxicology and Applied Pharmacology 172 (3) (May 1): 217–24. doi:10.1006/taap.2001.9156.</li>
</ul>
<ul>
<li>Norris DO. 2007. Vertebrate Endocrinology. Fourth ed. New York: Academic Press.</li>
</ul>
<ul>
<li>Reinsberg, Jochen, Petra Wegener-Toper, Katrin van der Ven, Hans van der Ven, and Dietrich Klingmueller. 2009. “Effect of Mono-(2-Ethylhexyl) Phthalate on Steroid Production of Human Granulosa Cells.” Toxicology and Applied Pharmacology 239 (1) (August 15): 116–23. doi:10.1016/j.taap.2009.05.022.</li>
</ul>
<ul>
<li>Rutherford R, Lister A, Hewitt LM, MacLatchy D. Effects of model aromatizable (17α-methyltestosterone) and non-aromatizable (5α-dihydrotestosterone) androgens on the adult mummichog (Fundulus heteroclitus) in a short-term reproductive endocrine bioassay. Comp Biochem Physiol C Toxicol Pharmacol. 2015 Apr;170:8-18. doi: 10.1016/j.cbpc.2015.01.004.</li>
<li>Sharpe RL, MacLatchy DL, Courtenay SC, Van Der Kraak GJ. Effects of a model androgen (methyl testosterone) and a model anti-androgen (cyproterone acetate) on reproductive endocrine endpoints in a short-term adult mummichog (Fundulus heteroclitus) bioassay. Aquat Toxicol. 2004 Apr 28;67(3):203-15.</li>
<li>Shoemaker JE, Gayen K, Garcia-Reyero N, Perkins EJ, Villeneuve DL, Liu L, et al. 2010. Fathead minnow steroidogenesis: in silico analyses reveals tradeoffs between nominal target efficacy and robustness to cross-talk. BMC systems biology 4: 89.</li>
<li>Skolness SY, Durhan EJ, Garcia-Reyero N, Jensen KM, Kahl MD, Makynen EA, et al. 2011. Effects of a short-term exposure to the fungicide prochloraz on endocrine function and gene expression in female fathead minnows (Pimephales promelas). Aquat Toxicol 103(3-4): 170-178.</li>
<li>Villeneuve DL, Breen M, Bencic DC, Cavallin JE, Jensen KM, Makynen EA, et al. 2013. Developing Predictive Approaches to Characterize Adaptive Responses of the Reproductive Endocrine Axis to Aromatase Inhibition: I. Data Generation in a Small Fish Model. Toxicological sciences : an official journal of the Society of Toxicology.</li>
<li>Villeneuve DL, Mueller ND, Martinovic D, Makynen EA, Kahl MD, Jensen KM, et al. 2009. Direct effects, compensation, and recovery in female fathead minnows exposed to a model aromatase inhibitor. Environ Health Perspect 117(4): 624-631.</li>
</ul>
<ul>
<li>Xu, Chuan, Ji-An Chen, Zhiqun Qiu, Qing Zhao, Jiaohua Luo, Lan Yang, Hui Zeng, et al. 2010. “Ovotoxicity and PPAR-Mediated Aromatase Downregulation in Female Sprague-Dawley Rats Following Combined Oral Exposure to Benzo[a]pyrene and Di-(2-Ethylhexyl) Phthalate.” Toxicology Letters 199 (3) (December 15): 323–32. doi:10.1016/j.toxlet.2010.09.015.</li>
</ul>
2016-11-29T18:41:332017-03-20T12:05:21bee27495-701a-44ac-8330-4372a7b5c15fd69edb12-dff7-43b0-8223-e3bca8b5a3e4<p>See Plausibility below.</p>
<p>Updated 2017-03-17.</p>
<p>Vitellogenin synthesis in fish is localized in the liver and is well documented to be regulated by estrogens via interaction with estrogen receptors (Tyler et al. 1996; Tyler and Sumpter 1996; Arukwe and Goksøyr 2003). The vitellogenin gene contains estrogen repsonsive elements in its promoter region and site directed mutagenesis has shown these to be essential for estrogen-dependent expression of vitellogenin (Chang et al. 1992; Teo et al. 1998). Liver is not regarded as a major site of E2 synthesis (Norris 2007), therefore the majority of E2 in liver comes from the circulation.</p>
<ul>
<li>Estrogen regulates expression of the vitellogenin gene in the amphibian Xenopus laevis (Skipper and Hamilton, 1977).</li>
</ul>
<h2> </h2>
<ul>
<li>Empirical support for estrogen-dependent regulation of vitellogenin synthesis:
<ul>
<li>Many studies have demonstrated that exposure of hepatocytes to estrogens in vitro or in vivo induce vitellogenin mRNA synthesis (e.g., see reviews by (Navas and Segner 2006; Iguchi et al. 2006)).</li>
<li>In female fathead minnows exposed to 17β-trenbolone, significant reductions in plasma E2 concentrations preceded significant reductions in plasma VTG (Ekman et al. 2011).</li>
<li>Intra-arterial injection of the estrogen 17α ethynyl estradiol into male rainbow trout causes vitellogenin induction with about a 12 h lag time before increasing from basal levels (Schultz et al. 2001).</li>
</ul>
</li>
<li>Specific empirical support for reductions in plasma E2 leading to reductions in hepatic vitellogenin synthesis:
<ul>
<li>In a number of time-course experiments with aromatase inhibitors (e.g., fadrozole, prochloraz), 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>It was demonstrated in Danio rerio that in vivo exposure to the aromatase inhibitor letrozole significantly reduced the expression of mRNA transcripts coding for vtg1, vtg2, and erα, all of which are known to be regulated by estrogens (Sun et al. 2010). However, similar effects were not observed in primary cultured hepatocytes from Danio rerio, indicating that letrozole’s effects on vtg transcription were not direct.</li>
<li> </li>
</ul>
</li>
</ul>
<p>Based on the limited set of studies available to date, there are no known inconsistencies.</p>
<ul>
<li>At least two computational models that include functions which link circulating concentrations of E2 to VTG production by the liver have been published (Li et al. 2011a; Murphy et al. 2005; Murphy et al. 2009), although both models focus on predicting plasma VTG concentrations rather than transcription or translation within the liver. A significant positive correlation (r=0.87) between plasma E2 concentrations corresponding plasma VTG concentrations in female fathead minnows held under laboratory conditions has also been reported (Ankley et al. 2008).</li>
</ul>
<ul>
<li>There are multiple isoforms of vitellogenin. The sensitivity and inducibility of each of those isoforms may vary somewhat. Consequently, response-response relationships may vary somewhat depending on the speicific isoform for which QPCR primers or antibodies were developed.</li>
</ul>
HighFemaleNot SpecifiedAdult, reproductively matureModerate<p>Key enzymes needed to synthesize 17β-estradiol first appear in the common ancestor of amphioxus and vertebrates (Baker 2011). However, non-oviparous vertebrates do not require vitellogenin. Consequently, this KER is applicable to oviparous vertebrates.</p>
<ul>
<li>Ankley GT, Bencic D, Cavallin JE, Jensen KM, Kahl MD, Makynen EA, et al. 2009b. Dynamic nature of alterations in the endocrine system of fathead minnows exposed to the fungicide prochloraz. Toxicol Sci 112(2): 344-353.</li>
<li>Ankley GT, Bencic DC, Cavallin JE, Jensen KM, Kahl MD, Makynen EA, et al. 2009a. Dynamic nature of alterations in the endocrine system of fathead minnows exposed to the fungicide prochloraz. Toxicological sciences : an official journal of the Society of Toxicology 112(2): 344-353.</li>
<li>Ankley GT, Miller DH, Jensen KM, Villeneuve DL, Martinovic D. 2008. Relationship of plasma sex steroid concentrations in female fathead minnows to reproductive success and population status. Aquatic toxicology 88(1): 69-74.</li>
<li>Arukwe A, Goksøyr A. 2003. Eggshell and egg yolk proteins in fish: hepatic proteins for the next generation: oogenetic, population, and evolutionary implications of endocrine disruption. Comparative Hepatology 2(4): 1-21.</li>
<li>Chang TC, Nardulli AM, Lew D, and Shapiro, DJ. 1992. The role of estrogen response elements in expression of the Xenopus laevis vitellogenin B1 gene. Molecular Endocrinology 6:3, 346-354\</li>
<li>Ekman DR, Villeneuve DL, Teng Q, Ralston-Hooper KJ, Martinovic-Weigelt D, Kahl MD, et al. 2011. Use of gene expression, biochemical and metabolite profiles to enhance exposure and effects assessment of the model androgen 17beta-trenbolone in fish. Environmental toxicology and chemistry / SETAC 30(2): 319-329.</li>
<li>Iguchi T, Irie F, Urushitani H, Tooi O, Kawashima Y, Roberts M, et al. 2006. Availability of in vitro vitellogenin assay for screening of estrogenic and anti-estrogenic activities of environmental chemicals. Environ Sci 13(3): 161-183.</li>
<li>Li Z, Kroll KJ, Jensen KM, Villeneuve DL, Ankley GT, Brian JV, et al. 2011a. A computational model of the hypothalamic: pituitary: gonadal axis in female fathead minnows (Pimephales promelas) exposed to 17alpha-ethynylestradiol and 17beta-trenbolone. BMC systems biology 5: 63.</li>
<li>Murphy CA, Rose KA, Rahman MS, 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. Environmental toxicology and chemistry / SETAC 28(6): 1288-1303.</li>
<li>Murphy CA, Rose KA, 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. Reproductive toxicology 19(3): 395-409.</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>Norris DO. 2007. Vertebrate Endocrinology. Fourth ed. New York: Academic Press.</li>
<li>Schultz IR, Orner G, Merdink JL, Skillman A. 2001. Dose-response relationships and pharmacokinetics of vitellogenin in rainbow trout after intravascular administration of 17alpha-ethynylestradiol. Aquatic toxicology 51(3): 305-318.</li>
<li>Skipper JK, Hamilton TH. 1977. Regulation by estrogen of the vitellogenin gene. Proc Natl Acad Sci USA 74:2384-2388.</li>
<li>Skolness SY, Durhan EJ, Garcia-Reyero N, Jensen KM, Kahl MD, Makynen EA, et al. 2011. Effects of a short-term exposure to the fungicide prochloraz on endocrine function and gene expression in female fathead minnows (Pimephales promelas). Aquat Toxicol 103(3-4): 170-178.</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(7): 793-799.</li>
<li>Teo BY, Tan NS, Lim EH, Lam TJ, Ding JL. A novel piscine vitellogenin gene: structural and functional analyses of estrogen-inducible promoter. Mol Cell Endocrinol. 1998 Nov 25;146(1-2):103-20. PubMed PMID: 10022768.</li>
<li>Tyler C, Sumpter J. 1996. Oocyte growth and development in teleosts. Reviews in Fish Biology and Fisheries 6: 287-318.</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>Villeneuve DL, Breen M, Bencic DC, Cavallin JE, Jensen KM, Makynen EA, et al. 2013. Developing Predictive Approaches to Characterize Adaptive Responses of the Reproductive Endocrine Axis to Aromatase Inhibition: I. Data Generation in a Small Fish Model. Toxicological sciences : an official journal of the Society of Toxicology.</li>
<li>Villeneuve DL, Mueller ND, Martinovic D, Makynen EA, Kahl MD, Jensen KM, et al. 2009. Direct effects, compensation, and recovery in female fathead minnows exposed to a model aromatase inhibitor. Environ Health Perspect 117(4): 624-631.</li>
</ul>
2016-11-29T18:41:332017-03-20T12:28:39e0a0f52c-0518-4181-93d4-6e7df48a229dc85210c6-86ac-48bd-965a-2828514873ac<p>SEE BIOLOGICAL PLAUSIBILITY BELOW</p>
<p>Updated 03/20/2017</p>
<p>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).</p>
<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 SpecifiedUnspecificNot SpecifiedAll life stagesModerate<p>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.</p>
<ul>
<li>Kidd KA, Blanchfield KH, Palace VP, Evans RE, Lazorchak JM, Flick RW. 2007. Collapse of a fish population after exposure to a synthetic estrogen. PNAS 104:8897-8901.</li>
<li>Kramer VJ, Etterson MA, Hecker M, Murphy CA, Roesijadi G, Spade DJ, Spromberg JA, Wang M, Ankley GT. Adverse outcome pathways and ecological risk assessment: bridging to population-level effects. Environ Toxicol Chem. 2011 Jan;30(1):64-76. doi: 10.1002/etc.375. PubMed PMID: 20963853</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. Ecotoxicology and Environmental Safety 59: 1-9.</li>
<li>Miller DH, Tietge JE, McMaster ME, Munkittrick KR, Xia X, Ankley GT. 2013. Assessment of Status of White Sucker (Catostomus Commersoni) Populations Exposed to Bleached Kraft Pulp Mill Effluent. Environmental toxicology and chemistry / SETAC (in press).</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 with zebrafish. (RIVM Report). Bilthoven, the Netherlands: Joint Dutch Environment Ministry.</li>
</ul>
2016-11-29T18:41:332017-03-20T13:49:05417da2f9-67a7-4892-ac8f-e54cf771a3c3e0a0f52c-0518-4181-93d4-6e7df48a229d<p>SEE BIOLOGICAL PLAUSIBILITY BELOW</p>
<p>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.</p>
<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>
<p>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.</p>
<p>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.</p>
HighFemaleHighAdult, reproductively matureModerateModerate<p>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.</p>
<ul>
<li>Ankley GT, Jensen KM, Makynen EA, Kahl MD, Korte JJ, Hornung MW, Henry TR, Denny JS, Leino RL, Wilson VS, Cardon MC, Hartig PC, Gray LE. Effects of the androgenic growth promoter 17-beta-trenbolone on fecundity and reproductive endocrinology of the fathead minnow. Environ Toxicol Chem. 2003 Jun;22(6):1350-60.</li>
<li>Eppig JJ. 1994. Further reflections on culture systems for the growth of oocytes in vitro. Human reproduction 9(6): 974-976.</li>
<li>Jamnongjit M, Hammes SR. 2005. Oocyte maturation: the coming of age of a germ cell. Seminars in reproductive medicine 23(3): 234-241.</li>
<li>Kang IJ, Yokota H, Oshima Y, Tsuruda Y, Shimasaki Y, Honjo T. The effects of methyltestosterone on the sexual development and reproduction of adult medaka (Oryzias latipes). Aquat Toxicol. 2008 Apr 8;87(1):37-46. doi: 10.1016/j.aquatox.2008.01.010.</li>
<li>Li Z, Villeneuve DL, Jensen KM, Ankley GT, Watanabe KH. 2011b. A computational model for asynchronous oocyte growth dynamics in a batch-spawning fish. Can J Fish Aquat Sci 68: 1528-1538.</li>
<li>Miller DH, Jensen KM, Villeneuve DL, Kahl MD, Makynen EA, Durhan EJ, et al. 2007. Linkage of biochemical responses to population-level effects: a case study with vitellogenin in the fathead minnow (Pimephales promelas). Environ Toxicol Chem 26(3): 521-527.</li>
<li>Tyler C, Sumpter J. 1996. Oocyte growth and development in teleosts. Reviews in Fish Biology and Fisheries 6: 287-318.</li>
</ul>
2016-11-29T18:41:332017-03-20T13:35:29a2ff8bda-aca7-4151-be8a-659d2e7d4cf0417da2f9-67a7-4892-ac8f-e54cf771a3c3<p>SEE BIOLOGICAL PLAUSIBILITY BELOW</p>
<p>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).</p>
<p>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.</p>
<p>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).</p>
<p>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.</p>
<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>
HighFemaleHighAdult, reproductively matureModerateModerate<p>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.</p>
<ul>
<li>Ankley GT, Jensen KM, Durhan EJ, Makynen EA, Butterworth BC, Kahl MD, et al. 2005. Effects of two fungicides with multiple modes of action on reproductive endocrine function in the fathead minnow (Pimephales promelas). Toxicol Sci 86(2): 300-308.</li>
<li>Ankley GT, Jensen KM, Makynen EA, Kahl MD, Korte JJ, Hornung MW, et al. 2003. Effects of the androgenic growth promoter 17-b-trenbolone on fecundity and reproductive endocrinology of the fathead minnow. Environmental Toxicology and Chemistry 22(6): 1350-1360.</li>
<li>Ankley GT, Kahl MD, Jensen KM, Hornung MW, Korte JJ, Makynen EA, et al. 2002. Evaluation of the aromatase inhibitor fadrozole in a short-term reproduction assay with the fathead minnow (Pimephales promelas). Toxicological Sciences 67: 121-130.</li>
<li>Arukwe A, Goksøyr A. 2003. Eggshell and egg yolk proteins in fish: hepatic proteins for the next generation: oogenetic, population, and evolutionary implications of endocrine disruption. Comparative Hepatology 2(4): 1-21.</li>
<li>Li Z, Villeneuve DL, Jensen KM, Ankley GT, Watanabe KH. 2011b. A computational model for asynchronous oocyte growth dynamics in a batch-spawning fish. Can J Fish Aquat Sci 68: 1528-1538.</li>
<li>Skolness SY, Blanksma CA, Cavallin JE, Churchill JJ, Durhan EJ, Jensen KM, et al. 2013. Propiconazole Inhibits Steroidogenesis and Reproduction in the Fathead Minnow (Pimephales promelas). Toxicological sciences : an official journal of the Society of Toxicology 132(2): 284-297.</li>
<li>Sun L, Zha J, Spear PA, Wang Z. 2007. Toxicity of the aromatase inhibitor letrozole to Japanese medaka (Oryzias latipes) eggs, larvae and breeding adults. Comp Biochem Physiol C Toxicol Pharmacol 145(4): 533-541.</li>
<li>Tyler C, Sumpter J. 1996. Oocyte growth and development in teleosts. Reviews in Fish Biology and Fisheries 6: 287-318.</li>
<li>Wallace RA, Selman K. 1981. Cellular and dynamic aspects of oocyte growth in teleosts. American Zoologist 21: 325-343.</li>
</ul>
2016-11-29T18:41:332017-03-20T13:21:09d69edb12-dff7-43b0-8223-e3bca8b5a3e4a2ff8bda-aca7-4151-be8a-659d2e7d4cf0<p>See biological plausibility, below.</p>
<p>Updated 03/20/2017.</p>
<p>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.</p>
<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>
<p>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.</p>
<p>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).</p>
Not SpecifiedUnspecificNot SpecifiedAdult, reproductively matureModerate<p>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).</p>
<ul>
<li>
<p><br />
</p>
</li>
<li>Ankley GT, Bencic D, Cavallin JE, Jensen KM, Kahl MD, Makynen EA, et al. 2009b. Dynamic nature of alterations in the endocrine system of fathead minnows exposed to the fungicide prochloraz. Toxicol Sci 112(2): 344-353.</li>
<li>Ankley GT, Bencic DC, Cavallin JE, Jensen KM, Kahl MD, Makynen EA, et al. 2009a. Dynamic nature of alterations in the endocrine system of fathead minnows exposed to the fungicide prochloraz. Toxicological sciences : an official journal of the Society of Toxicology 112(2): 344-353.</li>
<li>Ankley GT, Miller DH, Jensen KM, Villeneuve DL, Martinovic D. 2008. Relationship of plasma sex steroid concentrations in female fathead minnows to reproductive success and population status. Aquatic toxicology 88(1): 69-74.</li>
<li>Arukwe A, Goksøyr A. 2003. Eggshell and egg yolk proteins in fish: hepatic proteins for the next generation: oogenetic, population, and evolutionary implications of endocrine disruption. Comparative Hepatology 2(4): 1-21.</li>
<li>Bowman CJ, Kroll KJ, Hemmer MJ, Folmar LC, Denslow ND. 2000. Estrogen-induced vitellogenin mRNA and protein in sheepshead minnow (Cyprinodon variegatus). General and comparative endocrinology 120(3): 300-313.</li>
<li>Genovese G, Regueira M, Piazza Y, Towle DW, Maggese MC, Lo Nostro F. 2012. Time-course recovery of estrogen-responsive genes of a cichlid fish exposed to waterborne octylphenol. Aquatic toxicology 114-115: 1-13.</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>Li Z, Kroll KJ, Jensen KM, Villeneuve DL, Ankley GT, Brian JV, et al. 2011a. A computational model of the hypothalamic: pituitary: gonadal axis in female fathead minnows (Pimephales promelas) exposed to 17alpha-ethynylestradiol and 17beta-trenbolone. BMC systems biology 5: 63.</li>
<li>Murphy CA, Rose KA, Rahman MS, 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. Environmental toxicology and chemistry / SETAC 28(6): 1288-1303.</li>
<li>Murphy CA, Rose KA, 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. Reproductive toxicology 19(3): 395-409.</li>
<li>Schmid T, Gonzalez-Valero J, Rufli H, Dietrich DR. 2002. Determination of vitellogenin kinetics in male fathead minnows (Pimephales promelas). Toxicol Lett 131(1-2): 65-74.</li>
<li>Schultz IR, Orner G, Merdink JL, Skillman A. 2001. Dose-response relationships and pharmacokinetics of vitellogenin in rainbow trout after intravascular administration of 17alpha-ethynylestradiol. Aquatic toxicology 51(3): 305-318.</li>
<li>Skolness SY, Durhan EJ, Garcia-Reyero N, Jensen KM, Kahl MD, Makynen EA, et al. 2011. Effects of a short-term exposure to the fungicide prochloraz on endocrine function and gene expression in female fathead minnows (Pimephales promelas). Aquat Toxicol 103(3-4): 170-178.</li>
<li>Tyler C, Sumpter J. 1996. Oocyte growth and development in teleosts. Reviews in Fish Biology and Fisheries 6: 287-318.</li>
<li>Villeneuve DL, Breen M, Bencic DC, Cavallin JE, Jensen KM, Makynen EA, et al. 2013. Developing Predictive Approaches to Characterize Adaptive Responses of the Reproductive Endocrine Axis to Aromatase Inhibition: I. Data Generation in a Small Fish Model. Toxicological sciences : an official journal of the Society of Toxicology.</li>
<li>Villeneuve DL, Mueller ND, Martinovic D, Makynen EA, Kahl MD, Jensen KM, et al. 2009. Direct effects, compensation, and recovery in female fathead minnows exposed to a model aromatase inhibitor. Environ Health Perspect 117(4): 624-631.</li>
</ul>
2016-11-29T18:41:332017-03-20T12:58:16bee27495-701a-44ac-8330-4372a7b5c15fa2ff8bda-aca7-4151-be8a-659d2e7d4cf0<p>There is not a direct structural/functional relationship between reduced concentrations of 17ß-estradiol in plasma and reduced plasma VTG concentrations. The relationship is thought to be mediated through additional events of hepatic estrogen receptor activation, vitellogenin protein synthesis in the liver, and subsequent secretion of vitellogenin into the plasma. </p>
<p>Updated 2017-03-17</p>
<p>The mechanisms through which 17ß-estradiol stimulates the transcription and translation of hepatic vitellogenin are well understood.</p>
<ul>
<li>In fish, see: Tyler et al. 1996; Tyler and Sumpter 1996; Arukwe and Goksøyr 2003; Teo et al. 1998</li>
<li>In frogs: Chang et al. 1992; Wangh and Knowland 1975</li>
<li>In reptiles: Ho et al. 1980 </li>
<li>Ho (1987)</li>
<li>In birds: Deeley et al. 1975;</li>
</ul>
<p>17ß-estradiol is not synthesized in significant amounts in the liver. Its synthesis originates in other tissues, principally the gonads. It is then transported to the liver and other tissues via circulation (Norris 2007; Payne and Hales 2004; Miller 1988; Nagahama et al. 1993).</p>
<ul>
<li>Under conditions of continuous flow through exposure to 17ß-trenbolone (a non-aromatizable androgen receptor agonist), plasma E2 concentrations were reduced in female fathead minnows after 2, 4, or 8 d of exposure to concentrations of 0.05 ug/L or greater. Plasma VTG concentrations were significantly reduced only after 4 or 8 d of exposure, and at 4 d, only at a concentration of 0.5 ug/L, not 0.05 ug/L (Ekman et al. 2011).</li>
<li>In the same study by Ekman et al. (2011), once exposure ceased, plasma E2 concentrations returned to control levels within 48 h, while plasma VTG concentrations remained significantly depressed until d. 4, post-exposure.</li>
<li> Ankley et al. (2003) detected reductions in both plasma E2 and plasma VTG in female fathead minnows following 21 d of continuous exposure to 17ß-trenbolone. At 21 d, plasma E2 concentrations were impacted at concentrations of 0.5 ug/L or greater, while plasma VTG was significantly reduced at 0.05 ug/L or greater.</li>
<li>Villeneuve et al. (2016) observed significant reductions in both plasma E2 and plasma VTG in female fathead minnows exposed to 0.5 ug/L 17ß-trenbolone for 14 d.</li>
<li>Jensen et al. (2006) observed significant reductions in both plasma E2 and plasma VTG following exposure to 0.03 ug/L 17alpha-trenbolone for 21 d.</li>
<li>Following 21 d of continuous exposure to spironolactone, plasma E2 and plasma VTG were both significantly reduced in female fathead minnows. The lowest effect concentration for plasma E2 was 0.5 ug/L, while that for plasma VTG was 5 ug/L (LaLone et al. 2013).</li>
<li> In female Fundulus heteroclitus exposed to 5alpha-dihydrotestosterone for 14 d, plasma E2 was significantly reduced following exposure to 10 ug/L, while plasma VTG was reduced at 100 ug/L (Rutherford et al. 2015).</li>
<li>In two experiments in which female Fundulus heteroclitus were exposed to 17alpha-methyltestosterone, both plasma E2 and plasma VTG were significantly reduced. In both cases, plasma E2 was impacted at lower concentrations (0.25 ug/L in a 7 d study; 0.01 ug/L in a 14 d study) than plasma VTG (1 ug/L in the 7 d study; 0.1 ug/L in the 14 d study; Sharpe et al. 2004).</li>
<li>In two experiments where plasma E2 and plasma VTG were measured in female fathead minnows (Pimephales promelas) in a time-course following continuous exposure the aromatase inhibitor fadrozole, both plasma VTG and plasma E2 were depressed (Villeneuve et. al. 2009; 2013). In both cases, following cessation of exposure, plasma E2 concentrations recovered to control levels before plasma VTG concentrations recovered (Villeneuve et al. 2009; 2013).</li>
<li>Shroeder et al. (in preparation) reported effects on plasma E2 concentrations within 4 h of initiating exposure to 5 or 50 ug/L fadrozole. Plasma VTG concentrations did not decline until 24 h or later (Schroeder et al. 2009; Villeneuve et al. 2009; 2013).</li>
<li>In female fathead minnows exposed to 300 ug/L prochloraz, plasma E2 concentrations were significantly reduced after 12 h of exposure, while plasma VTG concentrations were not significantly reduced until 24 h of exposure (Skolness et al. 2011).</li>
<li>Ankley et al. (2009) reported significant reductions in plasma E2 in female fathead minnows following 24 h of exposure to 30 ug/L prochloraz. In the same study, plasma VTG concentrations did not significantly decline until 48 h of exposure, and then only at 300 ug/L prochloraz.</li>
<li>In a 21 d exposure to prochloraz, plasma E2 was significantly reduced in females exposed to 300 ug prochloraz/L, while plasma VTG was significantly reduced in females exposed to 100 ug/L (Ankley et al. 2005). </li>
</ul>
<ul>
<li>In several studies, significant decreases in plasma vitellogenin are detected at lower concentrations than those that result in significant decreases in plasma E2. However, detection of differences in plasma VTG is ofen enhanced by the greater dynamic range in the concentrations of the protein that occur in plasma, compared to the dynamic range of steroid hormone concentrations.</li>
</ul>
<ul>
<li>A computational model developed by Cheng et al. (2016) is capable of simulating altered plasma VTG concentrations associated with changes in plasma E2 concentrations in female fathead minnows. This model has been used to generate a quantitative response-response relationship that can predict steady state plasma VTG concentrations for a given steady state plasma E2 concentration (Conolly et al. 2017).
<ul>
<li>The model and response-response relationship were developed based on data from exposures to the model aromatase inhibitor fadrozole. The validity of the model-based predictions/relationships for other stressors and species has not yet been established.</li>
</ul>
</li>
<li>Li et al. (2011) also developed a physiologically-based computational model of the adult female fathead minnow (<em>Pimephales promelas</em>) hypothalamic-pituitary-gonadal axis. Conceptually, this model could also be applied to derive a quantitative response-response relationship between plasma E2 and plasma VTG concentrations. The Li et al. model was calibrated based on data from exposures to 17alpha-ethynylestradiol and 17ß-trenbolone. Neither its validity for other stressors or speices, nor its agreement with the Cheng et al. (2016) model have been examined in detail.</li>
</ul>
HighFemaleHighAdult, reproductively matureHighHigh<p>This key event relationship likely applies to oviparous vertebrates only.</p>
<ul>
<li>Key enzymes needed to synthesize 17β-estradiol first appear in the common ancestor of amphioxus and vertebrates (Baker 2011). </li>
<li>Vitellogenesis is common to a range of egg-laying vertebrates and invertebrates. However, in the case of invertebrates, vitellogenins are transported via hemolymph rather than plasma and vitellogenesis is regulated by invertebrate hormones, not estradiol.</li>
</ul>
<ul>
<li>Ankley GT, Jensen KM, Makynen EA, Kahl MD, Korte JJ, Hornung MW, Henry TR, Denny JS, Leino RL, Wilson VS, Cardon MC, Hartig PC, Gray LE. Effects of the androgenic growth promoter 17-beta-trenbolone on fecundity and reproductive endocrinology of the fathead minnow. Environ Toxicol Chem. 2003 Jun;22(6):1350-60.</li>
<li>Arukwe A, Goksøyr A. 2003. Eggshell and egg yolk proteins in fish: hepatic proteins for the next generation: oogenetic, population, and evolutionary implications of endocrine disruption. Comparative Hepatology 2(4): 1-21.</li>
<li>Baker ME. 2011. Origin and diversification of steroids: co-evolution of enzymes and nuclear receptors. Molecular and cellular endocrinology 334(1-2): 14-20.</li>
<li>Chang TC, Nardulli AM, Lew D, and Shapiro, DJ. 1992. The role of estrogen response elements in expression of the Xenopus laevis vitellogenin B1 gene. Molecular Endocrinology 6:3, 346-354\</li>
<li>Chang TC, Nardulli AM, Lew D, Shapiro DJ. The role of estrogen response elements in expression of the Xenopus laevis vitellogenin B1 gene. Mol Endocrinol. 1992 Mar;6(3):346-54. </li>
<li>Cheng WY, Zhang Q, Schroeder A, Villeneuve DL, Ankley GT, Conolly R. Computational Modeling of Plasma Vitellogenin Alterations in Response to Aromatase Inhibition in Fathead Minnows. Toxicol Sci. 2016 Nov;154(1):78-89.</li>
<li>Deeley RG, Mullinix DP, Wetekam W, Kronenberg HM, Meyers M, Eldridge JD, Goldberger RF. Vitellogenin synthesis in the avian liver. Vitellogenin is the precursor of the egg yolk phosphoproteins. J Biol Chem. 1975 Dec 10;250(23):9060-6.</li>
<li>Ekman DR, Villeneuve DL, Teng Q, Ralston-Hooper KJ, Martinović-Weigelt D, Kahl MD, Jensen KM, Durhan EJ, Makynen EA, Ankley GT, Collette TW. Use of gene expression, biochemical and metabolite profiles to enhance exposure and effects assessment of the model androgen 17β-trenbolone in fish. Environ Toxicol Chem. 2011 Feb;30(2):319-29. doi: 10.1002/etc.406.</li>
<li>Ho DM, L'Italien J, Callard IP. 1980. Studies on reptilian yolk:Chrysemys. Comp. Biochem. Physiol. 65B: 139-144.</li>
<li>Ho SM. Endocrinology of vitellogenesis. In Norris DO, Jones RE Eds, Hormones and reproduction in fishes, amphibians, and reptiles, Plenum, New York, (1987), pp. 146-169.</li>
<li>Jensen KM, Makynen EA, Kahl MD, Ankley GT. Effects of the feedlot contaminant 17alpha-trenbolone on reproductive endocrinology of the fathead minnow. Environ Sci Technol. 2006 May 1;40(9):3112-7.</li>
<li>LaLone CA, Villeneuve DL, Cavallin JE, Kahl MD, Durhan EJ, Makynen EA, Jensen KM, Stevens KE, Severson MN, Blanksma CA, Flynn KM, Hartig PC, Woodard JS, Berninger JP, Norberg-King TJ, Johnson RD, Ankley GT. Cross-species sensitivity to a novel androgen receptor agonist of potential environmental concern, spironolactone. Environ Toxicol Chem. 2013 Nov;32(11):2528-41. doi: 10.1002/etc.2330.</li>
<li>Li Z, Kroll KJ, Jensen KM, Villeneuve DL, Ankley GT, Brian JV, Sepúlveda MS, Orlando EF, Lazorchak JM, Kostich M, Armstrong B, Denslow ND, Watanabe KH. A computational model of the hypothalamic: pituitary: gonadal axis in female fathead minnows (Pimephales promelas) exposed to 17α-ethynylestradiol and 17β-trenbolone. BMC Syst Biol. 2011 May 5;5:63. doi: 10.1186/1752-0509-5-63.</li>
<li>Miller WL. 1988. Molecular biology of steroid hormone synthesis. Endocrine reviews 9(3): 295-318.</li>
<li>Nagahama Y, Yoshikumi M, Yamashita M, Sakai N, Tanaka M. 1993. Molecular endocrinology of oocyte growth and maturation in fish. Fish Physiology and Biochemistry 11: 3-14.</li>
<li>Norris DO. 2007. Vertebrate Endocrinology. Fourth ed. New York: Academic Press.</li>
<li>Payne AH, Hales DB. 2004. Overview of steroidogenic enzymes in the pathway from cholesterol to active steroid hormones. Endocrine reviews 25(6): 947-970.</li>
<li>Rutherford R, Lister A, Hewitt LM, MacLatchy D. Effects of model aromatizable (17α-methyltestosterone) and non-aromatizable (5α-dihydrotestosterone) androgens on the adult mummichog (Fundulus heteroclitus) in a short-term reproductive endocrine bioassay. Comp Biochem Physiol C Toxicol Pharmacol. 2015 Apr;170:8-18. doi: 10.1016/j.cbpc.2015.01.004.</li>
<li>Sharpe RL, MacLatchy DL, Courtenay SC, Van Der Kraak GJ. Effects of a model androgen (methyl testosterone) and a model anti-androgen (cyproterone acetate) on reproductive endocrine endpoints in a short-term adult mummichog (Fundulus heteroclitus) bioassay. Aquat Toxicol. 2004 Apr 28;67(3):203-15.</li>
<li>Teo BY, Tan NS, Lim EH, Lam TJ, Ding JL. A novel piscine vitellogenin gene: structural and functional analyses of estrogen-inducible promoter. Mol Cell</li>
<li>Tyler C, Sumpter J. 1996. Oocyte growth and development in teleosts. Reviews in Fish Biology and Fisheries 6: 287-318.</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>Villeneuve DL, Jensen KM, Cavallin JE, Durhan EJ, Garcia-Reyero N, Kahl MD, Leino RL, Makynen EA, Wehmas LC, Perkins EJ, Ankley GT. Effects of the antimicrobial contaminant triclocarban, and co-exposure with the androgen 17β-trenbolone, on reproductive function and ovarian transcriptome of the fathead minnow (Pimephales promelas). Environ Toxicol Chem. 2017 Jan;36(1):231-242. doi: 10.1002/etc.3531.</li>
<li>Wangh LJ, Knowland J. 1975. Synthesis of vitellogenin in cultures of male and female frog liver regulated by estradiol treatment in vitro. Proc. Nat. Acad. Sci. 72: 3172-3175.</li>
</ul>
2017-03-17T17:13:262018-10-18T11:02:14Aromatase inhibition leading to reproductive dysfunctionAromatase inhibition leading to reproductive dysfunction<p>Dan Villeneuve, US EPA Mid-Continent Ecology Division (villeneuve.dan@epa.gov)</p>
Open for citation & commentWPHA/WNT EndorsedIncluded in OECD Work Plan1.12<p>This adverse outcome pathway details the linkage between inhibition of gonadal aromatase activity in females and reproductive dysfunction, as measured through the adverse effect of reduced cumulative fecundity and spawning. Initial development of this AOP draws heavily on evidence collected using 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 2012). 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 aromatase, or in silico predictions of the ability to inhibit aromatase, as a means to identify chemicals with known potential to adversely affect fish populations and potentially other oviparous vertebrates.</p>
<p>Characterization of chemical properties: Chemicals are known to inhibit aromatase activity through two primary molecular mechanisms. Steroid-like structures can inhibit the enzyme at its active site, with structures having ∆4 positioned double bonds generally acting as stronger inhibitors than those with ∆5 positioned double bonds (Petkov et al. 2009). Non-steroidal aromatase inhibitors generally act by interfering with electron transfer via the cytochrome P450 heme group of the aromatase enzyme, with greater nucleophilicity of the heteroatom contributing to greater potency as an inhibitor (Petkov et al. 2009). Petkov et al. (Petkov et al. 2009) have provided a detailed analysis of structural categorization of chemicals as potential steroidal or non-steroidal aromatase inhibitors.</p>
<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>
adjacentModerateHighadjacentModerateHighadjacentModerateHighadjacentModerateModerateadjacentModerateModerateadjacentLowModerateadjacentModerateHighnon-adjacentModerateHigh<p>Support for the essentiality of a number of key events in the AOP was provided by several time-course, stop-reversibility, experiments with fathead minnows exposed to aromatase inhibitors.</p>
<p>1. Villeneuve et al. 2009 and 2013 examined a time-course of key event responses to fadrozole as well as the time-course of recovery following cessation of fadrozole delivery. Once fadrozole was removed from the system, ex vivo E2 production increased, followed by increases in plasma E2 concentrations, and then increases in plasma vitellogenin concentrations. Additionally, while exposure to the chemical was on-going, compensatory up-regulation of CYP19a1a gene expression resulted in increases in ex vivo E2 production, followed by increased plasma E2 and plasma VTG. The essentiality of aromatase inhibition relative to impaired E2 production was further supported by the observation of an "overshoot" in E2 production, relative to controls, shortly after cessation of fadrozole delivery.</p>
<p>2. Similar support was provided in a study by Ankley et al. (2009a). Cessation of prochloraz delivery resulted in rapid recovery of ex vivo E2 production and plasma E2 concentrations, with recovery of vitellogenin concentrations lagging slightly behind. Increased expression of cyp19a1a mRNA during the exposure period aligned with increased ex vivo E2 production, and increased plasma E2, compared to the first day of exposure.</p>
<p><br />
Rationale for essentiality calls:</p>
<p>• Aromatase, inhibition: [Strong] There is good evidence from stop/reversibility studies that ceasing delivery of the aromatase inhibitor leads to recovery of the subsequent key events.</p>
<p>• 17beta-estradiol synthesis by ovarian granulosa cells, reduction: [Strong] In both exposure studies and stop/reversibility studies, when ex vivo E2 production (as measure of this KE) recovers either through compensation or due to removal of the stressor, subsequent KEs have been shown to recover after a lag period.</p>
<p>• plasma 17beta-estradiol concentrations, reduction: [Strong] In both exposure studies and stop/reversibility studies, when plasma E2 concentrations recover either through compensation or due to removal of the stressor, subsequent KEs have been shown to recover after a lag period.</p>
<p>• vitellogenin production in liver (transcription, translation), reduction: [Moderate] This endpoint was not specifically examined in stop/reversibility studies with aromatase inhibitors, but biological plausibility provides strong support for the essentiality of this event.</p>
<p>• plasma vitellogenin concentrations, reduction: [Strong] Shown to recover in a predictable fashion consistent with the order of events in the AOP in stop/recovery studies.</p>
<p>• vitellogenin accumulation into oocytes and oocyte growth/development, reduction: [Weak] Some contradictory evidence regarding the essentiality of this event. No stop/reversibility studies have explicitly considered this key event.</p>
<p>• cumulative fecundity and spawning, reductions: [Moderate] By definition, some degree of spawning is required to maintain population.</p>
HighFemaleNot SpecifiedAdult, reproductively matureModerateModerateHigh<ul>
<li><strong>Sex</strong>: The AOP applies to females only. Males have relatively low gonadal aromatase expression and activity and the androgen 11-KT, rather than the estrogen E2 is a stronger driver of reproductive functions in males. That said, at least in fish, there is a potential autocrine and paracrine for estrogens synthesized in the brain in regulating reproductive behaviors. However, those potential effects are addressed through an alternative AOP that shares the MIE of aromatase inhibition.</li>
<li><strong>Life stages</strong>: 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><strong>Taxonomic</strong>: 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>
<p>Support for the essentiality of a number of key events in the AOP was provided by several time-course, stop-reversibility, experiments with fathead minnows exposed to aromatase inhibitors.</p>
<p>1. Villeneuve et al. 2009 and 2013 examined a time-course of key event responses to fadrozole as well as the time-course of recovery following cessation of fadrozole delivery. Once fadrozole was removed from the system, ex vivo E2 production increased, followed by increases in plasma E2 concentrations, and then increases in plasma vitellogenin concentrations. Additionally, while exposure to the chemical was on-going, compensatory up-regulation of CYP19a1a gene expression resulted in increases in ex vivo E2 production, followed by increased plasma E2 and plasma VTG. The essentiality of aromatase inhibition relative to impaired E2 production was further supported by the observation of an "overshoot" in E2 production, relative to controls, shortly after cessation of fadrozole delivery.</p>
<p>2. Similar support was provided in a study by Ankley et al. (2009a). Cessation of prochloraz delivery resulted in rapid recovery of ex vivo E2 production and plasma E2 concentrations, with recovery of vitellogenin concentrations lagging slightly behind. Increased expression of cyp19a1a mRNA during the exposure period aligned with increased ex vivo E2 production, and increased plasma E2, compared to the first day of exposure.</p>
<p><br />
Rationale for essentiality calls:</p>
<p>• Aromatase, inhibition: [Strong] There is good evidence from stop/reversibility studies that ceasing delivery of the aromatase inhibitor leads to recovery of the subsequent key events.</p>
<p>• 17beta-estradiol synthesis by ovarian granulosa cells, reduction: [Strong] In both exposure studies and stop/reversibility studies, when ex vivo E2 production (as measure of this KE) recovers either through compensation or due to removal of the stressor, subsequent KEs have been shown to recover after a lag period.</p>
<p>• plasma 17beta-estradiol concentrations, reduction: [Strong] In both exposure studies and stop/reversibility studies, when plasma E2 concentrations recover either through compensation or due to removal of the stressor, subsequent KEs have been shown to recover after a lag period.</p>
<p>• vitellogenin production in liver (transcription, translation), reduction: [Moderate] This endpoint was not specifically examined in stop/reversibility studies with aromatase inhibitors, but biological plausibility provides strong support for the essentiality of this event.</p>
<p>• plasma vitellogenin concentrations, reduction: [Strong] Shown to recover in a predictable fashion consistent with the order of events in the AOP in stop/recovery studies.</p>
<p>• vitellogenin accumulation into oocytes and oocyte growth/development, reduction: [Weak] Some contradictory evidence regarding the essentiality of this event. No stop/reversibility studies have explicitly considered this key event.</p>
<p>• cumulative fecundity and spawning, reductions: [Moderate] By definition, some degree of spawning is required to maintain population.</p>
<p><strong>Biological plausibility</strong>: Biological plausibility refers to the structural or functional relationship between the key events based on our fundamental understanding of "normal biology". In general, the biological plausibility and coherence linking aromatase inhibition through decreases in circulating concentrations of E2 is very solid. The biochemistry of steroidogenesis and the predominant role of the gonad in synthesis of the sex steroids is well established. Similarly, the role of E2 as the major regulator of hepatic vitellogenin production is widely documented in the literature. The direct link between reduced VTG concentrations in the plasma and reduced uptake into oocytes is highly plausible, as the plasma is the primary source of the VTG. However, the direct connection between reduced VTG uptake and impaired spawning/reduced cumulative fecundity is more tentative. It is not clear, for instance whether impaired VTG uptake limits oocyte growth and failure to reach a critical size in turn impairs physical or inter-cellular signaling processes that promote release of the oocyte from the surrounding follicles. In at least one experiment, oocytes with similar size to vitellogenic oocytes, but lacking histological staining characteristic of vitellogenic oocytes was observed (R. Johnson, personal communication). Regulation of oocyte maturation and spawning involves many factors other than vitellogenin accumulation (Clelland and Peng, 2009). At present, the link between reductions in circulating VTG concentrations and reduced cumulative fecundity are best supported by the correlation between those endpoints across multiple experiments, including those that impact VTG via other molecular initiating events (Miller et al. 2007).</p>
<p> </p>
<p><strong>Concordance of dose-response relationships</strong>: Dose response concordance considers the degree to which upstream events are shown to occur at test concentrations equal to or lower than those that cause significant effects on downstream key events, the underlying assumption being that all KEs can be measured with equal precision. There are a limited number of studies in which multiple key events were considered in the same study. These were considered the most useful for evaluating the concordance of dose-response relationships. In general, effects on downstream key events occurred at concentrations equal to or greater than those at which upstream events occurred (Concordance table: <a class="external autonumber" href="https://aopwiki.org/wiki/images/4/45/Aromatase_inhibition_dose-response_concordance_table_rev1.pdf" rel="nofollow" target="_blank">[1]</a>). However, there are exceptions. There are cases where no significant effects on estradiol synthesis by ovarian granulosa cells (ovary explants) were observed, but significant effects on plasma E2 or VTG concentrations were observed. Likewise, there are cases where impacts on plasma VTG were observed at concentrations lower than those reported to reduce plasma E2 concentrations. Based on knowledge of the studies in question, the apparent lack of concordance in some cases is driven by two primary factors. First, differences in the sensitivity and dynamic range of the measurements being made. Second, the effects of compensatory responses along the HPG axis. For instance, although ex vivo E2 production is rapidly affected by exposure to fadrozole, it is also a response that is more rapidly corrected through upregulation of aromatase transcripts (see Villeneuve et al. 2009), meaning that it recovers more quickly than plasma concentrations of E2 or plasma VTG concentrations. Thus, at certain time points, one can get an apparent effect on plasma E2 or T without a measurable impact on E2 production by the gonad tissue, because the upstream insult occurred earlier in time and was subsequently offset by a compensatory response, but the compensation has yet to propagate through the pathway. Sensitivity and dynamic range of the measurement methods is also an issue. Vitellogenin concentrations have a highly dynamic range and can change by orders of magnitude. Other endpoints like plasma steroids are regulated in a narrower range, making differences more difficult to distinguish statistically. Therefore, in our assessment, the deviations from concordance do not call the KERs into question.</p>
<p> </p>
<p>The concentration-dependence of the key event responses with regard to the concentration of aromatase inhibitor has been established in vitro and/or in vivo for nearly all key events in the AOP.</p>
<ol>
<li>Concentration-dependent aromatase inhibition: (Villeneuve et al. 2006; Ankley et al. 2005; M et al. 2004; AM et al. 2000; Shilling et al. 1999)</li>
<li>Concentration-dependent decreases in E2 production in vitro, ex vivo: (Ankley et al. 2002; Villeneuve et al. 2007; Villeneuve et al. 2009; Ankley et al. 2005; a Marca Pereira et al. 2011; Lee et al. 2006).</li>
<li>Concentration-dependent decreases in circulating E2 concentrations: (Ankley et al. 2002; Villeneuve et al. 2009; Ankley et al. 2005; Ankley et al. 2009a; GT et al. 2001)</li>
<li>Concentration-dependent decreases in vitellogenin mRNA expression: (Sun et al. 2010; Sun et al. 2011; Zhang et al. 2008)</li>
<li>Concentration-dependent decreases in circulating vitellogenin concentrations: (Ankley et al. 2002; Villeneuve et al. 2009; Ankley et al. 2005; Ankley et al. 2009a; Sun et al. 2007; GT et al. 2001; Ralston-Hooper et al. 2013)</li>
<li>Concentration-dependent reductions in VTG uptake into oocytes or impaired oocyte development: Concentration-dependence of these effects has not been well demonstrated. The effects, when seen, have typically been documented at the greatest exposure concentration tested, but concentration-dependence of the severity or frequency of the impact was not documented (e.g., (Ankley et al. 2002; Ankley et al. 2005; Sun et al. 2007)</li>
<li>Concentration-dependent reductions in cumulative fecundity: (Ankley et al. 2002; Ankley et al. 2005; Sun et al. 2007; Zhang et al. 2008)</li>
<li>Declining population trajectory: Modeled population trajectories show a concentration-dependent reduction in projected population size, however, those results are driven by the concentration-dependence of cumulative fecundity. Population-level effects have not been measured directly.</li>
</ol>
<p><br />
<strong>Temporal concordance</strong>: Temporal concordance refers to the degree to which the data support the hypothesized sequence of the key events; i.e., the effect on KE1 is observed before the effect on KE2, which is observed before the effect on KE3 and so on. Temporal concordance of the AOP from aromatase inhibition to decreased E2 production, decreased circulating E2, and decreased plasma VTG concentrations has been established (e.g., (Villeneuve et al. 2009; Ankley et al. 2009a; Skolness et al. 2011). Temporal concordance has not been established beyond that key event, in large part due to disconnect in the time-scales over which the events can be measured. For example, most small fish used in reproductive toxicity testing will can spawn anywhere from once daily to several days per week. Given the variability in daily spawning rates, it is neither practical nor effective to evaluate cumulative fecundity at a time scale shorter than roughly a week. Since the impacts at lower levels of biological organization can be detected within hours of exposure, lack of impact on cumulative fecundity before the other key events are impacted cannot be effectively measured. Overall, among those key events whose temporal concordance can reasonably be evaluated, the temporal profile observed is consistent with the AOP.</p>
<p><strong>Consistency</strong>: We are aware of no cases where the pattern of key events described was observed without also observing a significant impact on cumulative fecundity. The final adverse outcome is not specific to this AOP. Many of the key events included in this AOP overlap with AOPs linking other molecular initiating events to reproductive dysfunction in small fish.</p>
<p><strong>Uncertainties, inconsistencies, and data gaps</strong>: The current major uncertainty in this AOP is whether there is a direct biological linkage 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.</p>
<p>Assessment of quantitative understanding of the AOP:</p>
<p>At present, quantitative understanding of the AOP is approaching the point where an in vitro measurement of aromatase inhibition could be used as an input parameter into a series of coupled computational models that could generate quantitative predictions across multiple key events (e.g., circulating E2 concentrations, circulating VTG concentrations, predicted impacts on cumulative fecundity, and effects on population trajectories). A sequence of supporting models has been coupled together and predictions have been made for novel aromatase inhibitors (identified through high throughput in vitro screening), but those predictions have not yet been validated experimentally. The present models are also unable to account for pharmacokinetic considerations (e.g., adsorption, distribution, metabolism/biotransformation, and elimination) and have demonstrated only partial success in simulating compensatory/feedback responses to aromatase inhibition (e.g., (Breen et al. 2013).</p>
<ul>
<li>The present AOP can provide potential support for the use of alternatives to the fish short term reproduction assay as a screen for aromatase inhibitors.</li>
</ul>
<ul>
<li>The present AOP can serve as a foundation for tiered testing strategies and IATA related to risk assessments on chemicals identified as aromatase inhibitors.</li>
</ul>
<ul>
<li>The present AOP can be used to guide endpoint selection for effects-based monitoring studies at sites where aromatase inhibition has been identified as a relevant biological activity of interest (e.g., through bioeffects prediction or bioeffects surveillance approaches; see Schroeder et al. 2016).</li>
</ul>
<p>Schroeder, A. L., Ankley, G. T., Houck, K. A. and Villeneuve, D. L. (2016), Environmental surveillance and monitoring—The next frontiers for high-throughput toxicology. Environ Toxicol Chem, 35: 513–525. doi:10.1002/etc.3309</p>
<ul>
<li>A series of computational models aligned with this AOP (i.e., a quantitative AOP construct) can be applied to estimate in vivo bench-mark doses based on in vitro screening results. Case studies evaluating this application are under way.</li>
</ul>
Not Specified<p>1. OECD. 2012. Test No. 229: Fish Short Term Reproduction Assay. Paris, France:Organization for Economic Cooperation and Development.</p>
<p>2. Petkov PI, Temelkov S, Villeneuve DL, Ankley GT, Mekenyan OG. 2009. Mechanism-based categorization of aromatase inhibitors: a potential discovery and screening tool. SAR QSAR Environ Res 20(7-8): 657-678.</p>
<p>3. Lephart ED, Simpson ER. 1991. Assay of aromatase activity. Methods Enzymol 206: 477-483.</p>
<p>4. Letcher RJ, van Holsteijn I, Drenth H-J, Norstrom RJ, Bergman A, Safe S, et al. 1999. Cytotoxicity and aromatase (CYP19) activity modulation by organochlorines in human placental JEG-3 and JAR choriocarcinoma cells. Toxicology and applied pharmacology 160: 10-20.</p>
<p>5. Sanderson J, Seinen W, Giesy J, van den Berg M. 2000. 2-chloro-triazine herbicides induce aromatase (CYP19) activity in H295R human adrenocortical carcinoma cells: a novel mechanism for estrogenicity. Toxicological Sciences 54: 121-127.</p>
<p>6. Villeneuve DL, Knoebl I, Kahl MD, Jensen KM, Hammermeister DE, Greene KJ, et al. 2006. Relationship between brain and ovary aromatase activity and isoform-specific aromatase mRNA expression in the fathead minnow (Pimephales promelas). Aquat Toxicol 76(3-4): 353-368.</p>
<p>7. Ankley GT, Kahl MD, Jensen KM, Hornung MW, Korte JJ, Makynen EA, et al. 2002. Evaluation of the aromatase inhibitor fadrozole in a short-term reproduction assay with the fathead minnow (Pimephales promelas). Toxicological Sciences 67: 121-130.</p>
<p>8. Castro LF, Santos MM, Reis-Henriques MA. 2005. The genomic environment around the Aromatase gene: evolutionary insights. BMC evolutionary biology 5: 43.</p>
<p>9. Norris DO. 2007. Vertebrate Endocrinology. Fourth ed. New York: Academic Press.</p>
<p>10. Yaron Z. 1995. Endocrine control of gametogenesis and spawning induction in the carp. Aquaculture 129: 49-73.</p>
<p>11. Havelock JC, Rainey WE, Carr BR. 2004. Ovarian granulosa cell lines. Molecular and cellular endocrinology 228(1-2): 67-78.</p>
<p>12. Villeneuve DL, Ankley GT, Makynen EA, Blake LS, Greene KJ, Higley EB, et al. 2007. Comparison of fathead minnow ovary explant and H295R cell-based steroidogenesis assays for identifying endocrine-active chemicals. Ecotoxicol Environ Saf 68(1): 20-32.</p>
<p>13. McMaster ME MK, Jardine JJ, Robinson RD, Van Der Kraak GJ. 1995. Protocol for measuring in vitro steroid production by fish gonadal tissue. Canadian Technical Report of Fisheries and Aquatic Sciences 1961 1961: 1-78.</p>
<p>14. Ankley GT, Jensen KM, Kahl MD, Makynen EA, Blake LS, Greene KJ, et al. 2007. Ketoconazole in the fathead minnow (Pimephales promelas): reproductive toxicity and biological compensation. Environ Toxicol Chem 26(6): 1214-1223.</p>
<p>15. Villeneuve DL, Mueller ND, Martinovic D, Makynen EA, Kahl MD, Jensen KM, et al. 2009. Direct effects, compensation, and recovery in female fathead minnows exposed to a model aromatase inhibitor. Environ Health Perspect 117(4): 624-631.</p>
<p>16. Baker ME. 2011. Origin and diversification of steroids: co-evolution of enzymes and nuclear receptors. Molecular and cellular endocrinology 334(1-2): 14-20.</p>
<p>17. Jensen K, Korte J, Kahl M, Pasha M, Ankley G. 2001. Aspects of basic reproductive biology and endocrinology in the fathead minnow (Pimephales promelas). Comparative Biochemistry and Physiology Part C 128: 127-141.</p>
<p>18. 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.</p>
<p>19. 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.</p>
<p>20. 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.</p>
<p>21. 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.</p>
<p>22. 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.</p>
<p>23. Tyler C, Sumpter J. 1996. Oocyte growth and development in teleosts. Reviews in Fish Biology and Fisheries 6: 287-318.</p>
<p>24. 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.</p>
<p>25. 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.</p>
<p>26. Miller DH, Ankley GT. 2004. Modeling impacts on populations: fathead minnow (Pimephales promelas) exposure to the endocrine disruptor 17b-trenbolone as a case study. Ecotoxicology and Environmental Safety 59: 1-9.</p>
<p>27. Ankley GT, Jensen KM, Durhan EJ, Makynen EA, Butterworth BC, Kahl MD, et al. 2005. Effects of two fungicides with multiple modes of action on reproductive endocrine function in the fathead minnow (Pimephales promelas). Toxicol Sci 86(2): 300-308.</p>
<p>28. Ankley GT, Bencic D, Cavallin JE, Jensen KM, Kahl MD, Makynen EA, et al. 2009a. Dynamic nature of alterations in the endocrine system of fathead minnows exposed to the fungicide prochloraz. Toxicol Sci 112(2): 344-353.</p>
<p>29. Skolness SY, Durhan EJ, Garcia-Reyero N, Jensen KM, Kahl MD, Makynen EA, et al. 2011. Effects of a short-term exposure to the fungicide prochloraz on endocrine function and gene expression in female fathead minnows (Pimephales promelas). Aquat Toxicol 103(3-4): 170-178.</p>
<p>30. Breen M, Villeneuve DL, Ankley GT, Bencic DC, Breen MS, Watanabe KH, et al. 2013. Developing Predictive Approaches to Characterize Adaptive Responses of the Reproductive Endocrine Axis to Aromatase Inhibition: II. Computational Modeling. Toxicological sciences : an official journal of the Society of Toxicology.</p>
<p>31. Breen MS, Villeneuve DL, Breen M, Ankley GT, Conolly RB. 2007. Mechanistic computational model of ovarian steroidogenesis to predict biochemical responses to endocrine active compounds. Annals of biomedical engineering 35(6): 970-981.</p>
<p>32. Shoemaker JE, Gayen K, Garcia-Reyero N, Perkins EJ, Villeneuve DL, Liu L, et al. 2010. Fathead minnow steroidogenesis: in silico analyses reveals tradeoffs between nominal target efficacy and robustness to cross-talk. BMC systems biology 4: 89.</p>
<p>33. Quignot N, Bois FY. 2013. A computational model to predict rat ovarian steroid secretion from in vitro experiments with endocrine disruptors. PloS one 8(1): e53891.</p>
<p>34. Ankley GT, Bencic DC, Cavallin JE, Jensen KM, Kahl MD, Makynen EA, et al. 2009b. Dynamic nature of alterations in the endocrine system of fathead minnows exposed to the fungicide prochloraz. Toxicological sciences : an official journal of the Society of Toxicology 112(2): 344-353.</p>
<p>35. Villeneuve DL, Breen M, Bencic DC, Cavallin JE, Jensen KM, Makynen EA, et al. 2013. Developing Predictive Approaches to Characterize Adaptive Responses of the Reproductive Endocrine Axis to Aromatase Inhibition: I. Data Generation in a Small Fish Model. Toxicological sciences : an official journal of the Society of Toxicology.</p>
<p>36. Ankley GT, Cavallin JE, Durhan EJ, Jensen KM, Kahl MD, Makynen EA, et al. 2012. A time-course analysis of effects of the steroidogenesis inhibitor ketoconazole on components of the hypothalamic-pituitary-gonadal axis of fathead minnows. Aquatic toxicology 114-115: 88-95.</p>
<p>37. Li Z, Kroll KJ, Jensen KM, Villeneuve DL, Ankley GT, Brian JV, et al. 2011a. A computational model of the hypothalamic: pituitary: gonadal axis in female fathead minnows (Pimephales promelas) exposed to 17alpha-ethynylestradiol and 17beta-trenbolone. BMC systems biology 5: 63.</p>
<p>38. A A, A G. 2003. Eggshell and egg yolk proteins in fish: hepatic proteins for the next generation: oogenetic, population, and evolutionary implications of endocrine disruption. Comparative Hepatology 2(4): 1-21.</p>
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