## WHITE PAPER

AQUATIC LIFE CRITERIA FOR CONTAMINANTS OF

EMERGING CONCERN

PART I

GENERAL CHALLENGES AND RECOMMENDATIONS

Prepared by the OW/ORD Emerging Contaminants Workgroup

June 03, 2008

NOTICE

THIS DOCUMENT IS AN INTERN AL PLANNING DOCUMENT

It has been prepared for the purpose of Research & Development Planning. It has not been formally released by the U.S. Environmental Protection Agency and should not at this stage be construed to represent Agency guidance or policy.ii EPA WORKGROUP

U.S. EPA, NHEERL, Mid-Continen t Ecology Division, Duluth, MN Gerald T. Ankley*

Richard Bennett

Russell J. Erickson* Dale J. Hoff, Workgroup Co-chair*

David R. Mount* Joseph Tietge

U.S. EPA, NHEERL, Gulf Ecolog y Division, Gulf Breeze, FL Geraldine Cripe

U.S. EPA, NERL, Ecological Exposure Research Division, Cincinnati, OH

Mitchell Kostich

David Lattier

James Lazorchak*

U.S. EPA, Office of Water, Washington, DC Janita Aguirre Joseph Beaman, Workgroup Co-chair* Diana Eignor

Lisa Huff

U.S. EPA, Office of Pesticide Programs, Washington, DC

Les Touart

Jean Holmes

Technical Support – Great Lakes En vironmental Center, Columbus, OH Tyler K. Linton* Gregory J. Smith

*Coauthoriii TABLE OF CONTENTS

1.0 INTRODUCTION ............................................................................................................... 1 1.1 What is a Contaminant of Emerging Concern? ................................................................1 1.2 Why is EPA Concerned with CECs?................................................................................2 1.3 Purpose and Organization of This White Paper................................................................3

2.0 CURRENT ALC METHODOLOGY..................................................................................5

2.1 Standard ALC Derivation Procedures ..............................................................................5

2.2 Alternatives for ALC Derivation ......................................................................................6 2.3 Precedent for Deviating from Ba sic ALC Derivation Procedures....................................9

3.0 CHARACTERISTICS OF CECs AND THEIR INFLUENCE ON ALC

DEVELOPMENT..............................................................................................................10

3.1 Characteristics of HPG-Active EDCs.............................................................................10 3.2 Implications for ALC Development ...............................................................................12 3.2.1 The Need For and Relevance of Acute Toxicity Data and a CMC ................... 13 3.2.2 Defining Minimum Data Requirements in Terms of Taxonomic Coverage ..... 14 3.2.3 Defining Appropriate Chronic Toxicity Data.................................................... 16 3.2.4 Selecting Effect Endpoints Upon Which to Base ALC ..................................... 18 3.2.4.1 Specific Examples of Measurable Cha nges at Different Levels of Biological Organization....................................................................................................... 20 3.3 Pathways and Receptors Beyond the HPG-Axis ............................................................25

4.0 SUMMARY AND RECOMMENDATIONS....................................................................27 4.1 Relevance of Acute Toxicity Effect Levels in Setting ALC for CECs...........................27 4.2 Defining Minimum Data Requirement s in Terms of Taxonomic Coverage ..................28 4.3 Use of Non-Resident Speci es in ALC Development......................................................29 4.4 Defining Appropriate Chronic Toxicity Data.................................................................29 4.5 Selection of Effect(s) Endpoints Upon Which to Base ALC..........................................30 4.6 Involvement of an Expert Panel......................................................................................31

5.0 REFERENCES ................................................................................................................. .32iv List of Acronyms:

ACR Acute to Chronic Ratio ALC Aquatic Life Criteria Ah Aryl Hydrocarbon (receptor) AV Acute (toxicity) value AWQC Ambient Water Quality Criteria

CCC Criterion Continuous Concentration CEC Contaminants of Emerging Concern

CMC Criterion Maximum Concentration CV Chronic (toxicity) value

CWA Clean Water Act CYP Cytochrome enzymes (P450) EDC Endocrine Disrupting Chemicals E2 Estradiol (natural estrogen) EE2 Ethynylestradiol (synthetic pharmaceutical estrogen) ELS Early Life-Stage (toxicity test) EPA Environmental Protection Agency

FAV Final Acute Value

FACR Final Acute to Chronic Ratio

GMAV Genus Mean Acute Value

GMCV Genus Mean Chronic Value HPG Hypothalamic-Pituitary-Gonadal (axis) HPT Hypothalamic-Pituitary-Thyroid (axis)

LOEC Lowest Observed Effect Concentration MDR Minimum Data Requirement

MOA Mode of Action

NOEC No Observed Effect Concentration OECD Organization for Economic Development and Cooperation OWC Organic Wastewater Contaminant PBDE Polybrominated Diphenyl Ether

PPCP Pharmaceutical and Personal Care Product POP Persistent Organic Pollutant SMAV Species Mean Acute Value TBT Tributyltin USGS United States Geological Survey VTG Vitellogenin protein

vtg Vitellogenin gene transcript

WWTP Wastewater Treatment Plant1 1.0 INTRODUCTION

Under the United States Clean Water Act (C WA) (33 U.S.C. Sections 1251-1387), EPA is required to take a number of acti ons to protect and restore the eco logical integrity of the Nation’s water bodies. Under Section 304(a) of the CWA, EPA must deve lop and publish ambient water quality criteria. Ambient water quality criteria (A WQC) are levels of individual pollutants, water quality characteristics, or descriptions of cond itions of a water body that, if met, should protect the designated use(s) of the water. Examples of designated uses of a water body include swimming, drinking water, fishing, fish spawning, and navigation. St ates and authorized tribes establish designated uses for th eir water bodies. AWQC are reco mmended guidance that states and tribes may use as part of their water qual ity standards to protec t water bodies for their designated use from chemical pollutants.

AWQC for aquatic life (aquatic life criteria, ALC) developed under Section 304(a) reflect the “latest scientific knowledge” conc erning “all identifiable effects” of the pollutant in question. These criteria are based solely on data and scient ific determinations on the relationship between

environmental concentrations of the pollutant and its effects. Cr iteria do not consider social and economic impacts, or the technological feasibility of meeting the chemical concentration values in ambient water. Since the early 1980's, EP A has been developing ALC to protect aquatic organisms from chemical specific pollutants under Section 304(a) of the CWA. In 1985, EPA published Guidelines for Deriving Numerical National Water Quality Criteria for the Protection of Aquatic Organisms and Their Uses (hereafter referred to as the “ Guidelines ”; Stephan et al. 1985). The Guidelines has provided uniformity and transp arency in the derivation methodology of ALC for a large number of compounds among seve ral classes of chemical s. The majority of EPA’s currently recommended ALC have been derived using the methods outlined in the

Guidelines

.

While the Guidelines remain the primary instrument the Agen cy uses to meet its broad objectives for the development of ALC, there have been ma ny advances in aquatic sciences, aquatic and wildlife toxicology, population mo deling, and ecological risk asse ssment that are relevant to deriving ALC. Some of the advances have b een addressed through supplemental guidance on the derivation or site-specific modification of criteri a (Prothro 1993; U.S. EPA 1994a), while others have been incorporated directly into derivati on of individual ALC for certain chemicals (e.g., saltwater chronic aquatic life criterion for tributyltin, U.S. EPA 2003). Recently, considerable attention has been generated by a widely rangi ng group of chemicals termed, in this document, contaminants of emerging concern (CECs). As is discussed in the body of this document, some of these CECs present challenges for the application of the Guidelines to ALC development.

1.1 What is a Contaminant of Emerging Concern?

The term “contaminant of emerging concern” is be ing used within the Office of Water to replace “emerging contaminant,” a term that has been used loosely since the mid-1990s by EPA and others to identify chemicals and other substances that have no regulator y standard, have been recently “discovered” in natural streams (oft en because of improved analytical chemistry detection levels), and potentially cause deleteri ous effects in aquatic life at environmentally2 relevant concentrations. They are pollutants not currently included in routine monitoring programs and may be candidates for future regula tion depending on their ( eco)toxicity, potential health effects, public perception, and frequency of occurrence in environmental media. CECs are not necessarily new chemicals. They include po llutants that have often been present in the environment, but whose presence and si gnificance are only now being evaluated.

CECs include several types of chemicals:

• Persistent organic pollutants (POPs) such as polybrominated diphenyl ethers (PBDEs; used in flame retardants, furniture foam , plastics, etc.) and other global organic contaminants such as perfluorinated organic acids;

• Pharmaceuticals and personal care products ( PPCPs), including a wide suite of human prescribed drugs (e.g., antidepressants, blood pressure), over-the -counter medications (e.g., ibuprofen), bactericides (e.g., tric losan), sunscreens, synthetic musks;

• Veterinary medicines such as antimicrobials , antibiotics, anti-funga ls, growth promoters and hormones;

• Endocrine-disrupting chemi cals (EDCs), including s ynthetic estrogens (e.g.,17 α

ethynylestradiol, which also is a PCPP) and androgens (e.g., trenbo lone, a veterinary drug), naturally occurring estrogens (e.g.,17ß-es tradiol, testosterone), as well as many others (e.g., organochlorine pesticides, al kylphenols) capable of modulating normal hormonal functions and steroidal sy nthesis in aquatic organisms;

• Nanomaterials such as carbon nanotubes or na no-scale particulate ti tanium dioxide, of

which little is known about either thei r environmental fate or effects.

1.2 Why is EPA Concerned with CECs?

The variety of chemicals labeled as CECs leads to a variety of concerns for EPA. Widespread uses, some indication of chemical persistence, effects found in natural systems, and public

concerns have made clear the need for EPA to de velop criteria that can be used to help assess and manage potential risk of some CECs in th e aquatic environment. A recent U.S. Geological Survey (USGS) reconnaissance study (Kolpin et al. 2002) pr ovides a good example of the prevalence of a wide range of CECs in U.S. st reams. Improved analytical chemistry techniques

were used to document the occurrence of wh at the authors called organic wastewater contaminants (OWCs) being released into surf ace waters from wastewater treatment plants (WWTPs). The targeted OWCs included PPCPs, veterinary medicines and other EDCs. The investigators found at least one of 95 different target OWCs in 80 percent of the 139 streams sampled. A median of seven, and as many as 38, OWCs were found in single samples.

The use and occurrence patterns as sociated with CECs are varied. Some CECs are similar to conventional toxic pollutants in that they are a ssociated with industria l releases, whereas many others are used by the general public every day in homes, on farms, by businesses and industry (Daughton and Ternes 1999). PPCPs acting as EDCs can be released direc tly to the environment after passing through wastewater treatment pro cesses, which are typically not designed to remove these pollutants from the effluent (Hallin g-Sorensen et al. 1998). Sludge from secondary treatment processes are land-applied as biosolid s, supplying CECs which may leach or run off into nearby bodies of water. Pharmaceuticals used in animal feeding operations may be released3 to the environment in animal wastes via dir ect discharge of aquacu lture products (i.e., antibiotics), the excretion of substances in anim al urine and feces of livestock animals, and the washoff of topical treatments from livestock animals (Boxall et al. 2003).

EDCs discharged at WWTPs are one group of CECs with potentially widespread environmental effects (Folmar et al. 1996; Folmar et al. 2001 ; Jobling et al. 1998; Woodling et al. 2006). Although particular concern has been expressed about the anthropogenic EDCs, there are also natural estrogens (estradiol and its metabolites estriol and estrone) entering the aquatic environment through wastewater discharge and excr etion from domestic animals. Furthermore,

little is known about the envir onmental occurrence, fate and, transport for any of these compounds after they enter aquatic ecosystems. Many of the man-made compounds have been in use for a long time, and there is concern about pharmacologically active ingredients and personal care products that are designed to stimulate a physiological response in humans, plants, and animals (Daughton and Ternes 1999).

Frequent detection of compounds by itself does not constitute a need for ALC. Rather, criteria development for CECs needs to focus efforts on chemicals that demonstrate a reasonable potential to adversely affect aquatic life. Of CECs now known to be found in some surface waters of the U.S., EDCs have received the most attention because field studies from around the world have demonstrated that very low con centrations of some of these compounds can significantly impact natural popula tions of aquatic vertebrates. For example, observational field studies (Jobling et al. 1998) have shown a high occurrence of inters ex (the presence of both male and female characteristics) in wild populations of a fish known as roach ( Rutilus rutilus ) in rivers in the United Kingdom that are downstream from WWTPs. Similar results have recently been reported for white sucker ( Catastomus commersoni ) in northern Colorado, U.S.A (Woodling et al. 2006). In a multiyear study by Kidd et al. (2007) , the authors showed that environmentally

relevant concentrations of et hynylestradiol, EE2, caused reproductive failure and near collapse of a natural fathead minnow population in an experime ntal lake, and also had deleterious effects on the reproductive biology of the pearl dace. These direct effects resu lting in loss of forage fish have led to cascading effects on the lake trout population due to lack of prey (Kidd, personal communication). Researchers from the U.S. Geol ogical Survey (USGS) have observed intersex and testis-ova (the presence of eggs in the testis) in bass spec ies collected from the Potomac River and its tributaries in West Virginia, Maryland, and Washington DC, and also quantified EDCs in their blood (Blazer et al. 2007; Chambers and Leiker 2006). The occurrence of intersex fish in the Potomac River, as well as documented occurrence of this and re lated effects in other waters of the US and internationally, prompted C ongressional hearings that were held in October 2006 to inquire about the “State of the Science on EDCs in the Environment,” as well as EPA

activities associated with EDCs.

1.3 Purpose and Organization of This White Paper

The purpose of this white paper is to provide ge neral guidance on how cr iteria development for CECs could be facilitated through a supplemental interpretation of the Guidelines , with particular attention to PPCPs with an EDC mode of action (MOA). Section 2 of this part (Part I) of the white paper describes the Guidelines procedures and identifies se veral areas in which these4 procedures could be modified to address poten tial limitations for deriving criteria for CECs. Section 3 expands upon the areas of concern with respect to specific toxi cological characteristics

of some CECs. Section 4 summa rizes these concerns and provi des recommendations that could aid in the development of crite ria for CECs in a resource efficient manner that takes best advantage of existing knowledge. Part II of this white paper further describes these concerns and recommendations using data for the synthetic pharmaceutical estrogen EE2.5 2.0 CURRENT ALC METHODOLOGY

The Guidelines specify various data and procedural recommendations for criteria derivation, and also define general risk management goals for criteria, which are to provide a high level of protection for aquatic communities and for important species in these communities. ALC are

defined to consist of two concentrations – the Criterion Maximum Concentration (CMC), intended to protect against seve re acute effects, and the Crit erion Continuous Concentration (CCC), intended to protect against longer term e ffects on survival, growth, and reproduction. The

CMC is used in criteria to lim it peak exposures by requiring th at 1-hour averages of exposure

concentrations not exceed the CMC more often th an once in three years on average. The CCC is

used in criteria to limit more prolonged exposures by requiring th at 4-day averages of exposure

concentrations not exceed the CCC more ofte n than once in three years on average.

The CMC and CCC are usually derived from laborat ory toxicity test re sults using specific standard procedures described in the Guidelines , but the Guidelines also have general provisions for deviating from these procedures as warrant ed by available information. The following text will first give an overview of the data requirement s and calculations in the standard procedures, and then discuss how these procedures might vary under the umbrella of the Guidelines

.

2.1 Standard ALC Derivation Procedures

The CMC is determined based on available Acute Values (AVs) – median lethal concentrations ( LC 50 s) or median effect (for a severe acute e ffect such as immobilization) concentrations ( EC 50 s) from aquatic animal acute toxicity tests (48- to 96-hours long) meeting certain data quality requirements. To compute a CMC, the Guidelines require that acceptable AVs be available for at least eight genera with a specifi ed taxonomic diversity, in order to address a wide variety of the organisms constituting an a quatic animal community. These minimum data requirements include three vertebrates (a salmon id fish, a fish from a family other than salmonidae, and a species from a third chordate family) and five invertebrates (a planktonic crustacean, a benthic crustacean, an insect, a species from a phylum other than Chordata or Arthropoda, and a species from another or der of insect or a fourth phylum).

For each genus, a Genus Mean Acute Value (GMAV) is calculated by first taking the geometric mean of the available AVs within each species (Species Mean Acute Value, SMAV) and then the geometric mean of the SMAVs within the genus. The fifth percentile of the set of GMAVs so obtained is calculated based on a specified es timation method, and designated the Final Acute Value (FAV). The FAV is then lowered to the SMAV for an important, sensitive species if appropriate. The CMC is set equal to half of the FAV to represent a low level of effect for the fifth percentile genus, rather than 50% effect.

The CCC is generally determined based on availabl e Chronic Values (CVs), which are either (a) the geometric mean of the highest no-observe d-effect concentration (NOEC) and lowest6 observed effect concentration (LOEC) for effect s on survival, growth, or reproduction in aquatic animal chronic tests or (b) in some recent criteria (e.g., ammonia), the EC 20 in such tests based on concentration-effect regression analyses. Chroni c tests for invertebrate species are required to include the entire life-cycle, but for fish partial life-cycle and early life-stage (ELS) testing protocols are accepted, the latter not including re productive endpoints and not used if life-cycle or partial life-cycle tests are available a nd show more sensitive adverse effects.

If CVs are available for at least eight genera with the required taxonom ic diversity, the CCC is set to the fifth percentile of Ge nus Mean Chronic Values (GMCVs), by the same procedure used

to derive an FAV from GMAVs. Otherwise, the CCC is set to the FAV divided by a Final Acute Chronic Ratio” (FACR) that is based on acute to chronic ratios (ACRs – the ratio of the AV to the CV from parallel acute and ch ronic tests) for at least three species with a specified taxonomic diversity. The CCC can also be based on plant toxici ty data if aquatic plan ts are more sensitive than aquatic animals, or on other data as deemed scientifically justified.

Further details on test require ments and calculation methods fo r the CMC and CCC are specified

in the Guidelines , including deriving criteria that are a function of water quality characteristics.

2.2 Alternatives for ALC Derivation

The procedures described above enable broad app lication to toxic chemicals generally, and are only constrained by specific data requir ements for quality and minimum taxonomic representation. Since they are not restricted with respect to specific types of chemicals, there is no reason to suppose that the standard data requirements and procedures specified by the

Guidelines are any more or less applicable to CECs th an to the chemicals for which criteria have already been developed. The Guidelines anticipated that rote applica tion of the basic procedures may not yield the most appropria te criteria; consequently, the Guidelines provide flexibility when appropriate for deviation from the normal procedures regardless of the type of chemical, as indicated by the following provisions (hereafter referred to as the "Good Science" clauses:

“These National Guidelines should be modified whenever sound scientific evidence

indicates that a national cr iterion produced using these Guidelines would probably be substantially overprotective or underprotec tive of aquatic organisms and their uses on a

national basis.” (p. 18).

"On the basis of all available pertinent l aboratory and field inform ation, determine if the criterion is consistent with sound scientific evidence. If it is not, another criterion, either higher or lower, should be derived using appr opriate modifications of these Guidelines." (p. 57).

In addition, although the sta ndard procedures in the Guidelines for deriving a CMC and CCC use only toxicity tests meeting certain requirements, the Guidelines also mandate the collation and examination of other data that might show effects that should be considered in criteria derivation:7 "Pertinent information that c ould not be used in earlier sections might be available concerning adverse effects on aquatic organism s and their uses. The most important of these are data on ... any other adverse effect that has been show n to be biologically important. Especially important are data for species for which no other data are available. ... Such data might affect a criterion if th e data were obtained with an important species, the test concentrati ons were measured, and the endpoint was biologically important.” (p. 54).

While alternatives are allowed when a specific situation dictates, the Guidelines still require that any changes in the procedures are consistent w ith the level of protec tion represented by the standard procedures:

"Derivation of numerical national water qual ity criteria for aquatic organisms and their uses is a complex process and requires knowle dge in many areas of aquatic toxicology; any deviation from these Guidelines should be carefully considered to ensure that it is consistent with other parts of these Guidelines." (p. iv).

Therefore, for the development of criteria for an y chemical, the general strategy should be to

start with the standard Guidelines procedures and then to adapt t hose procedures as warranted by

available information on the effects of the chem ical. This strategy applies to CECs as well, although certain considerations might more cons istently be important for CECs. Specific attributes of CECs that might aff ect criteria derivations are considered in Section 3 of this paper,

but several issues are introduced he re that are of general concern.

Are data on acute toxicity needed for risk assessments? Some chemicals are not acutely t oxic even at concentrations so high that they could not possibly occur in the environment (e.g., at the chemical solubility, or exceeding exposures possible based on known chemical production and discharges). The acu te lethality of some classes of chemicals might be measurable, but would occur at enviro nmental concentrations so much higher than those affecting reproduction, grow th, or chronic survival that , in practice, environmental exposures will always be far below acutely lethal levels if those exposures are managed to limit chronic effects. Therefore, derivation of the CM C might be unnecessary or impossible. Thus, if the existing data indicate that it is reasonably certain that acute toxicity would not occur at environmentally relevant concentrations, conduc ting additional acute te sts is likely to be

unwarranted.

Even if a CMC is not needed, another use of acu te toxicity data is for developing “acute to chronic ratios” (ACRs) that are used with the FAV to calculate the CCC (see pages 40-42 in the

Guidelines ) , so that dropping acute test ing requirements must consider this consequence as well. However, if acute effect concentrations ar e extremely high compared to chronic effect concentrations (large ACRs), whether the ACR approach should be even used warrants some consideration. Large ACRs are not, per se , less accurate than low ACRs, provided acute and

chronic effect concentrations are well defined and the issue is si mply extrapolating from acute to chronic toxicity within a species. However, for cr iteria calculations, the FACR needs to be a ratio that extrapolates from the fifth percentile of the acute effect concentration distribution to the fifth percentile of the chronic eff ect distribution. This requires appropriately combining ACR8 information across species, the accuracy of which might be affected by large ACRs even if the accuracy of the individual ACRs is not. Therefor e, in addition to not needing a CMC, extreme acute tolerance might also warrant direct calc ulation of the CCC rath er than using the ACR approach, and thus eliminate the need for fulf illing all of the minimum acute toxicity test requirements as specified by the Guidelines

.

How should data requirements fo r tolerant taxa be addressed? The fifth percentile calculation methods for th e CMC (as well as the CCC if the eight minimum data requirements noted above are met) require actual GMAV (or GMCV) values only for the four most sensitive genera. For more tolerant ge nera, it is only necessary to know that these toxicity values are greater than t hose of the four sensitive species. Therefore, toxicity test results that report "greater than" effect concentrations are acceptable for the tolerant taxa, and in fact are used in various criteria already.

If chronic tests have not al ready been done on some taxa needed for the minimum data requirements, but which are known to be tolera nt, testing resources might be wasted by generating numbers that will not aff ect results. If methods such as inter-chemical or inter-species extrapolation methods, or assays (e.g., in vitro tests, biomarkers) that have been related to apical effects such as reductions in growth, survival, or reproduction can demonstrate these taxa to be insensitive compared to other taxa , actual chronic tests on these ta xa may not be needed. In other words, can minimum data requirements for tolerant taxa be satisfied by some type of estimation rather than by an actual test result?

However, adding estimated data can become a rather open-ended process. Therefore, consideration must be given to how many estimated values should be allowed, relative to measured values, to produce an appropriate distribu tion of taxa in the data set used for criteria

derivation.

Should fish chronic tests be required to address reproduction? For chemicals (e.g., environmental estrogens) for wh ich reproductive toxicity is of most concern,

the allowance in the Guidelines for using ELS tests might need reconsideration. The Guidelines already give priority to life-cyc le and partial life-cycle tests wh en they are available and show greater sensitivity than ELS tests. However, other information (from other species, similar chemicals, knowledge of the MOA) regarding la tent or multigenerati onal reproductive effects may demonstrate the importance of sexual develo pment and reproduction, so as to establish a basis for not considering ELS test results (or ev en partial life-cycle test s), but rather requiring life-cycle tests for fish.

What endpoints can serve as surrogates for traditional chronic endpoints? Although chronic criteria are and will continue to be based on effects on reproduction, growth, and survival, another issue is whether only toxicity data directly addressing these endpoints is acceptable. Is there additional information (e.g., sub-organismal biomarkers, behavioral data)

that can be used in criteria derivation because they are adequately correlated to reproduction, growth, and survival? Use of such da ta would be consistent with the Guidelines requirements to examine all pertinent data and make modifications to the criteria derivatio n procedures that are

consistent with sound scientific evidence9 2.3 Precedent for Deviating from Basic ALC Derivation Procedures

The recent ALC document for tribut yltin (TBT) provides a good exampl e of some of the types of procedural criteria modifications discussed above . TBT is a highly toxic biocide that has been used extensively in anti-fouling pa int to protect the hulls of large ocean-going ships. It is deemed a problem in the aquatic environment because it is extremely toxic to non-target organisms, and has been linked to imposex (the superimposition of male anatomical characteristics on females) and to immuno-supression in snails and bivalv es (U.S. EPA 2003). The concentrations reported to cause imposex in the laboratory are lo wer (range: 0.0093 to 0.0334 μg/L) than the FCV (0.0658 μg/L) calculated using the standard ALC derivation procedures (U.S. EPA 2003). The

low effect concentrations established for female gastropods in the laboratory were subsequently

corroborated in field studies. The CCC was lowe red (to 0.0074 μg/L) based on the judgment that

these effects were relevant for the risks of TBT to gastropod reproduction.10 3.0 CHARACTERISTICS OF CECS AND THEIR INFLUENCE ON ALC

DEVELOPMENT

As described in Section 1.0, chemicals become labe led as CECs for a variety of reasons, many of which have relatively little to do with their t oxicological characteristic s. Consequently, the Guidelines cannot be interpreted or modified in one particular way that would be universally appropriate for all CECs. However, some charact eristics may be shared by various CECs, such that discussing the implications of these characteristics in the context of deriving water quality criteria is worthwhile. The expected outcome is additional guidance addressing key issues that may arise and how best to accommodate these issues in deriving criteria.

Much of the technical discussion that follows is centered on EDCs and, ev en more specifically, around chemicals that interact with the hypotha lamic/pituitary/gonadal (HPG) axis. Endocrine

function controlled via the HPG axis involves hormones broadly known as estrogens (“female” hormones such as estradiol) and androgens (“ma le” hormones such as testosterone), along with the body tissues and biochemical machinery with whic h they interact. Effects on this part of the endocrine system of various aquatic species have been documented in the literature, and publicized in the media, making toxicological disruption of this mechanism a good choice for discussing CECs in the context of the Guidelines. However, these types of substances are only a subset of EDCs, and an even smaller subset of CECs as a whole. While much of the discussion that follows uses HPG-active chemicals as a poin t of reference, the concepts presented may be

useful in the derivation of ALC fo r many other chemicals as well. It is the principles more than the specifics that are important in considering the content of this report.

3.1 Characteristics of HPG-Active EDCs

While estrogenic and androgenic hormones are a co re component of the HPG axis, this system also includes a much larger group of tissues a nd biochemical machinery within the body which, in vertebrates, govern sexual development, maturation, and re production. Commensurate with this complexity, there are many places within th e system that environmental chemicals may act to modify the normal function of the HPG-ax is. Thinking simply of “estrogenicity” or “androgenicity” as toxicological modes of action is still too broad – these categorical classes are more the outward “symptoms” of disruption in the HPG axis than they are unique modes of toxicological action. For example, the synthetic st eroids EE2 and trenbolone bind to (and act as agonists of) vertebrate estrogen and androgen receptors, respectiv ely, with similar or greater affinity than the natural endogenous hormones, estr adiol and testosterone. In contrast, a variety of other medicinal pharmaceuticals are specificall y designed to do the opposite, to be antagonists of these same receptors. As examples, tamoxifen (breast cancer treatment) and flutamide (prostate cancer treatment) bind quite specifically to vertebrate estrogen and androgen receptors, respectively, thereby blocking the activity of endogenous steroid hormones. But disruptors of the

HPG axis are not limited to chemicals that bind directly to estrogen or androgen receptors; they

also include chemicals that interact elsewh ere in the overall biochemical pathway. As an example, there are chemicals that exert their act ivity through interactions with CYP (cytochrome11 P450) enzymes involved in ster oid production. The pharmaceutical chemical fadrozole acts to inhibit CYP19 aromatase, the enzyme that conve rts estradiol to testos terone. A number of conazole fungicides act as competitive inhibitors of several CYPs further up the steroid synthesis pathway (Ankley et al. 2005).

Unlike many other chemicals that have either n on-specific (e.g., narcotics) or more generalized reactive modes of action (e.g., electrophilic chem icals interacting with nucleic acids and proteins), HPG-active compounds te nd to have very specific in teractions with particular molecular targets within the biochemical pathway. There are a number of consequences arising from this specificity. One important conseque nce of target specificity is potency. Many pharmaceuticals are designed to be highly specific, and thus are extremely potent. For example, EE2 and trenbolone affect reproduction and developm ent in fish at water concentrations in the very low ng/L (part-per-trillion) range (e.g., Ankl ey et al. 2003; Länge et al. 2001), well below

effect concentrations for most chemicals for wh ich current ALC have been derived. These very low biologically-active concentrations pres ent substantial challenges for analytical

determinations associated either with lab- based effects testing or field monitoring of in situ exposures. In the case of EE2 a nd/or trenbolone, effects observed in fish are at concentration levels below the methodological limit of detection for most laboratories even in laboratory test water, and even more so ambient water and effluents.

Such high potency can influence how one would a pproach criteria derivation when the chemicals exert minimal acute toxicity, but cause mostly long-term, sub-leth al effects. Trenbolone and EE2 illustrate this situation quite well. Like most pharmaceuticals (some exceptions being chemotherapy and anti-parasitical agents), these chemicals are designed to “adjust” the biochemistry of the body without causing acute t oxicity or other signif icantly adverse side effects. As a consequence, these types of pharmaceu ticals tend to have low toxicity in short-term lethality assays (Webb 2001). In the context of cr iteria development, this has implications for the

use of ACRs. Most conventiona l toxic pollutants with EPA AL C have ACRs of 10 or less (Cunningham et al. 2006; Host et al. 1995). In contrast, ACRs for EE2 and 17 β -trenbolone in fish have been shown to range from 1,000 to greater than 300,000 (Ankley et al. 2005). Again, this is a result of relatively low acute toxicity and high chronic potency. Importantly, limited data for other MOA classes of pharmaceuticals suggest that this phenomenon is not restricted to

endocrine-active substances. For ex ample, Huggett et al. (2002) repor ted an ACR in fish of about 50,000 for propanolol, a β -blocker. As discussed in Section 2, this large difference in acute and chronic potency may both make CMC values moot in the environmental management of these chemicals, and introduce uncertainty in the extr apolation between acute and chronic effects in

the derivation of a CCC.

The specificity of the molecular target also can gr eatly affect those taxa likely to be sensitive to

the chemical MOA of concern. While some bi ological pathways (e.g., energy metabolism) tend to be highly conserved across a ll organisms, others can be quit e specific to certain phylogenetic groups. Although the control of reproductive functi on through the HPG axis is highly conserved across vertebrate classes, lower taxonomic groups such as inverteb rates have different endocrine system structure that function differently. For example, Segner et al. (2003) tested several estrogenic chemicals, including EE2, in a variety of partial and full life-cycle tests with a model fish (zebrafish) and several aquatic invertebrate species. They found that the fish was by far most12 sensitive to the effects of the estrogenic chemi cals, and was the only species that responded to EE2 at environmentally-relevant concentrations. As a result, it is likely that chronic toxicity data

for fish would be the most influential in se tting the criterion for EE2, and correspondingly unlikely that toxicological data for invertebrate species would have much impact. Plants do not have comparable endocrine system structure or function, and would not be expected to be sensitive to these types of compound s, but there is research that indicates that algal species may be a uniquely sensitive taxa for the assessment of other types of CECs such as antibacterial products like triclosan (Orvos et al. 2002; Wilson et al. 2003).

Specificity in MOA can also affect how or if effe cts are expressed within a toxicity test, even in potentially sensitive species. In the case of ch emicals that affect endoc rine function, there are distinct “windows” when animals are likely to be sensitive and/or exhibi t adverse outcomes. For example, a popular amphibian early developmen tal assay-FETAX (Frog Embryo Teratogenesis Assay- Xenopus ) would be inadequate for detecting thyr oid-active toxicants because the period of exposure and observation occurs early in developm ent before the thyroid axis is functional in Xenopus . In the case of HPG-active toxicants, there are two window s of sensitivity during an animal’s life: during sexual differentiation in developing organisms (when “organizational” alterations occur), and during ac tive reproduction in adults (when “activational” responses can be manifested; Ankley and Johnson 2004). As a result, it is critical that testing for HPG-active EDCs occur during periods when the system is vulne rable to disruption. It is equally critical that toxicity tests include observati ons during the periods when eff ects are expressed. Some of the changes caused during sensitive early developmental windows may not be expressed until later in life. For example, the ELS toxicity test prot ocol commonly used in criteria development to estimate the chronic sensitivity of fish contains th e early life stages that could be sensitive to disruption of sexual development, but it does not extend through to maturation, and would

therefore be insensitive for detecti ng disruption of sexual development.

3.2 Implications for ALC Development

As is clear from the text above, some charac teristics of HPG-active chemicals (and many other CECs) create the need to carefully interpret the intent of the Guidelines , not just the routine derivation process. As indicated in Section 2.0 there is nothing a bout CECs that invalidates the principles embodied in the Guidelines ; however, the Guidelines were written before many of the issues discussed in Section 3.1 were known, so they do not necessarily contain prescriptive guidance for some of the nuances created by some CECs. The following paragraphs discuss the implications of these issues for criteria deve lopment, following the four general topic areas

outlined in Section 2.0:

• The need for and relevance of acu te toxicity data and a CMC;

• Defining minimum data requirements in terms of taxonomic coverage;

• Defining appropriate chr onic toxicity data; and

• Selecting effect endpoints upon which to base criteria13 3.2.1 The Need For and Relevance of Acute Toxicity Data and a CMC

As described in Section 2.2, th ere are two primary uses for acute toxicity data under the

Guidelines : 1) the derivation of the CMC; and 2) establishment of the CCC when the FAV/FACR method is used. As a practical matter, if the CMC is more than 96-fold higher than the CCC, then it will always be the CCC that is more limiting. This is because in the standard formulation of criteria, the CMC has a one -hour averaging peri od and the CCC a 96-hour averaging period; thus, if the di fference between them is more than 96-fold, it is mathematically impossible to exceed the CMC without also exceed ing the CCC. A minor exception to this issue occurs when ALC are implemented in an NPDES permit such that the CMC is applied to whole effluent while the CCC is applied after mixing, a nd the available in-strea m dilution is large. However, these exceptions are rare, and even th e 96-fold difference discussed here pales in comparison to the factors of 1,000 to 300,000-fold di scussed previously in regard to EE2 and trenbolone. In cases where such extreme differences between acute and chronic toxicity thresholds exist, establishing ALC as ha ving only a CCC seems a reasonable approach.

While it is easy to see why a CMC would not be necessary when you have sufficient acute test

data to show that the CMC would be dramatically highe r than the CCC, this begs the question of

how much data are needed to decide that this is the case. This decision should occur during the “problem formulation” step in the risk assessment for a specific chemical/class, and should be guided by the following types of information:

• the amount and phylogenetic spread of acute toxicity data th at are available;

• toxicity data from short-term exposures that do not meet the strict definitions in the

Guidelines of acute toxicity data acceptable for criteria derivation, but from which information on responses to acute exposures can be inferred;

• short-term effect data garner ed from longer-term exposures;

• information from closely related chemicals th at are thought to have the same MOA, and

have more robust acu te data sets; and

• knowledge of the degree of phylogenetic distri bution of the toxicity pathway of concern.

A complicating issue resulting from a “moot” CMC is that data availability for acute effects will likely be limited. As such, having less than the required acute MDRs may preclude the ability to derive a CCC using the FAV/FACR approach typically used in the Guidelines . For chemicals with highly specific modes of action and large ACRs (such as many EDCs), it is very likely that the mechanisms for acute and chronic toxicity diffe r, since biological activity resulting in chronic effects is designed into the pr oduct and not a secondary conse quence - such as many of the historical contaminants for which EPA has develope d criteria. Also, sensitiv ity of different taxa

classes to acute and/or chronic t oxicity varies widely due to pres ence (or absence) and structure and function of phylogenetically-conserved system s. Both of these issues would introduce considerable uncertainty into the availability and interpretation of ACRs, and probably make it inadvisable to use the FAV/FACR approach anyway. The Guidelines discuss the inadvisability of using the ACR approach when ACRs vary by more than a factor of 10 without a clear relationship to taxonomy or acute sensitivity (page 41 of the Guidelines ). A more advisable approach would generally be to develop a CCC dire ctly from a sufficiently robust set of chronic data, using the procedures outlined in the Guidelines or an appropriate modification thereof.14 3.2.2 Defining Minimum Data Requirement s in Terms of Taxonomic Coverage

To develop a CCC directly from chronic toxi city data (rather than via FAV/FACR), the

Guidelines require that acceptable chronic toxicity data be available from at least eight families with a taxonomic distribution fulfilling the requirements specified in the Guidelines (referred to as the “minimum data requirements” or “MDRs”). Having a blanket requirement for meeting the eight MDRs was included to insure a minimum level of “certainty” that the Guidelines will be protective of the broad phylogenetic distribut ion of organisms found in aquatic systems. Including this phylogenetic spread also enables criteria to be de veloped for chemicals for which the toxicological MOA is not known. Instead of “ knowing” what organisms are most likely to be sensitive to a particular chemi cal, requiring a broad spread of em pirical toxicity data makes it likely that whatever taxa may be sensitive to a ch emical, they will be represented to some degree in the toxicity data set.

In the case of EDCs, PPCPs, and certain other chemical classes, we may have a reasonable understanding of the toxicological MOA for the chemical, and from that knowledge we may be able to infer what taxa are most likely to be se nsitive to a particular chemical (Ankley et al. 2007; Williams 2005). As discussed in Section 2.2, the procedure used in the Guidelines for estimating the 5 th percentile of a toxicity distribution is dependent on only the four lowest values; for this reason exact values for insensitive genera are not necessary, as long as there are sufficient data to infer that their sensitivity is lower than the four most sensitive genera.

So how does one determine that a particular taxon is insensitive? The ge neral structure of the

Guidelines presumes that sensitivity is determined by conducting an acceptable toxicity test with that taxon. However, one can infer th at the actual need is only to have sufficient information to

conclude that the taxon is insens itive; if that determination ca n be confidently made based on other information, the information need may be met even if an actual toxicity test with that particular organism and chemical has not been conducted. This does not change the intent of the

Guidelines . It only acknowledges the possi bility that there is more than one way to meet the information requirement.

Using the example of EE2, there is both physiological understandi ng and some empirical toxicity data to support the belief that vertebrates will be far more sensitive to EE2 toxicity than will invertebrates (see Part II of this white paper and Segner et al. 2003). As such, it would seem inappropriate to invest resources in testing a wide range of invertebrate taxa classes for sensitivity when all existing data indicate that the data would not affect the final criterion, which would be driven by sensitivity of vertebrates. In this case, it makes sense to argue that certain of the eight MDRs might be declared met not th rough direct testing, but through toxicological understanding of the chemical’s MOA and the physio logy of those other taxa or existing toxicity data that establishes sensitivity differences among taxa.

While this logic is clear, one must be careful in presuming that the primary MOA demonstrated by organisms with the target physiology is the on ly toxic MOA for the chemical. Particularly given the phylogenetic diversity of organisms, it is always possible that a chemical that behaves15 with one MOA in one class of organisms may ex ert toxicity through a different mechanism in a different phylogenetic group. There are precedents for this scenario (Ankley et al. 2007). For example, exposure to the non-steroidal anti-inf lammatory drug diclofenac via consumption of dead livestock has greatly diminished some populations of vultures in several East Asian

countries. Diclofenac kills the bi rds through renal failure, which is only a relatively minor sideeffect of the drug in mammals. While the mech anism of renal toxicity in vultures is likely molecularly related to the mechanism of therapeu tic action in man, i.e., both seem to occur due to inhibition of prostaglandin s ynthesis (Meteyer et al . 2005), the inhibition of similar molecular components appears to be manifested as drama tically different whole organism endpoints. The key is in achieving a reasonabl e balance between expending res ources on collecting data most likely to influence the criterion, while mainta ining some kind of backstop against initially unexpected toxicity in other organisms.

One possibility for enhancing confidence regard ing phylogenetic sensitivity is in considering data for other, closely-related chem icals with the same MOA. While the Guidelines focus analysis on toxicity data for th e specific chemical in question, an understanding of toxicological

MOA can also lead to an unders tanding of how other chemicals might act to exploit the same biological system in the same way. For example, one might reasonably in fer that the relative species sensitivity to EE2 is likely similar to 17- ß estradiol (E2), the natural hormone which EE2 mimics. If, for example, there were a taxon whic h had been tested and found insensitive to E2, but had not been tested with EE2, it seems a reas onable assumption that that taxon would also be

insensitive to EE2.

The possibility of fulfilling certain information requirements using data other than from direct toxicity testing does raise some other interpretati on challenges, in particular the definition of the sample size for determining the 5 th percentile of the genera sens itivity distribution. For example,

if one has reason to believe that all crustace ans would be insensitive to a chemical, how many genera does that assertio n represent in the calcul ation of the genera sensitivity distribution?

While this is a real issue that will have to be addressed, we believe the problem is tractable and

the details of the resolution are left to later work.

Because of the risk that our mechanistic under standing of a chemical may be incomplete, it seems unlikely that one could justify complete ly bypassing several MDRs solely on theoretical arguments (e.g., developing a criterion for a testos terone mimic based only on chronic toxicity data for vertebrates, with no invertebrate data at all). At the same time , prudent application of other data types to fulfill certain information requirements for criteria derivation does seem appropriate. Given the tremendous va riation in understand ing and availability of data likely to exist for different CECs, it is presumed that at le ast initial application of this approach will have to be justified on a chemical-by- chemical basis using appropriate scientific judgment. However, lines of evidence that might be applicable to this determination include:

• an in-depth understanding of the toxicologi cal (or, in the case of drugs, therapeutic) MOA;

• information on the basic physiology of ot her taxa in relation to the MOA;

• toxicity data from chronic exposures or othe r relevant experiments that do not meet the strict definitions of acceptable chronic data given in the Guidelines , but from which16 information on relative taxon sensitivity can be inferred; and

• information from closely related chemicals that are thought to have the same MOA and

have more robust acute or chronic data sets.

A separate, but related issue arises in respect to data from species not resident to North America.

The Guidelines specify the use of toxicity data only from species resident to North America. However, particularly in regard to the study of EDCs, some fish sp ecies not resident to the U.S. have been advanced as experimental models for the evaluation of the chronic effects of EDCs to fish. Two clear examples at the time of this report are the zebrafish ( Brachydanio rerio ) and the Japanese medaka ( Oryzias latipes ), for which equivalency of ED C test data (with the fathead minnow) has been proposed through international groups such as the Organization for Economic Cooperation and Development (OECD; Ankley and Johnson 2004). These species have a rich toxicological database, and we know of no reason to believe that their sensitivities would be expected to be substantively different from sensitivit ies of at least some fish resident to the U.S. In keeping with international ha rmonization, we suggest that toxi city data from species with recognized international equivalenc y be included in cr iteria derivation with the full weight given to data from resident species.

3.2.3 Defining Appropriate Chronic Toxicity Data

The Guidelines state that acceptable chronic tests for criteria derivation are full life-cycle exposures (F 0 egg to F 1 offspring) for both vert ebrates and invertebrates, as well as partial lifecycle (adult to juvenile) and early life-stage (ELS; egg to juvenile) tests for fish. The acceptance of ELS tests in particular as acceptable chronic te sts is predicated on the work of McKim et al. (1978) and other evidence that the toxicity th resholds obtained from ELS tests are generally

within a factor of two of the thresholds from life-cycle chronic tests.

While this general approach has been applied with apparent success for many chemicals, the

Guidelines intimate concerns with the approach, noting that for some chemicals, ELS tests might not be good predictors of chronic toxicity, whic h would violate the principle underlying the use of ELS tests as chronic data (page 39 in the Guidelines ). As noted previously, toxicological data

for chemicals like EE2 show that certain chemical s may have potent effect s on life processes that lie outside the exposure period represente d by ELS tests (e.g., pronounced effects on reproduction), or on life processes for which the expression of effects doe s not occur until after the ELS period (e.g., embryo or la rval exposure resulting in e ffects on sexual development and maturation in adult fish; see Secti on 3.1). It is clear from these ex amples that there are chemicals

for which ELS tests should not be used as surr ogates for full life-cycle exposures. In fact, chemicals that affect sexual differentiation may no t be adequately assessed even with partial lifecycle exposures, since these protocols do not generally include observation of sexual development/maturation in fish e xposed during early development.

While the “Good Science” clause and other text in the Guidelines would not support reliance on

ELS tests as chronic data for chemicals known to have specific effects on other life processes, such as sexual developm ent or reproduction, the Guidelines would allow the development of a criterion using only ELS data for fish if there we re not any specific data to indicate that this approach would be inappropriate. This is akin to an “innocent until proven guilty” approach.17 However, we believe experiences with chemicals like EDCs make clear the need to move from the previous approach to one of “guilty until proven innocent.” In other words, it is probably wiser to require that the chronic toxicity data for fish include exposur e and observation over a full life-cycle unless there is an affirmative reason to believe that it is not necessary (note: this issue is equally relevant to invertebrates species, but the ELS tests discussed in the Guidelines are focused explicitly on fish; invertebrate tests would already be required to be life-cycle). In keeping with this shift in emphasis, we believe the requirements for chronic toxicity data in the

Guidelines should be tightened by adding the fu rther requirements that either:

1) Full life-cycle data be availabl e for at least one fish species; or

2) There is a body of experimental informa tion indicating that li fe processes outside the ELS or partial life-cycle expos ure/observation windows would not be important to capturing the important t oxicological effects of the chemical.

At first glance, #2 may seem like requiring the pr oof of a negative, in that one would have to actually conduct the life-cy cle test required by #1 in order to show that #2 is true. However, we believe there may be circumstances in which there may be data that speak to the sensitivity of different life stages that come fr om studies that, while scientifi cally valid, for some reason do not meet all the requirements of a valid life-cycle test as defined in the Guidelines . For example, there may be data for a life-cycle test with a non -resident species that includes the relevant life processes but does not qualify as an acceptable chr onic test for the derivati on of criteria because it is non-resident. Alternatively, there may be data from experiments that violate other requirements of acceptable toxicity data under the Guidelines , but still provide insight into sensitive exposure periods or life processes. Even though CVs from such data may not be used directly to calculate a chronic criterion, it seems reasonable to us e such data to evaluate the question of where in the life-cy cle there are important windows of exposure and/or effect and how that constrains the adequacy of ELS tests to represent chronic toxicity. In other cases, there may be sufficient information from other types of research to de monstrate to a reasonable level of certainty that a chemical’s toxicological mechanism(s) would not preclude the use of ELS tests as indicators of chronic toxicity.

Where life-cycle toxicity data are available, the results of those experime nts should be carefully examined to determine the likelihood that impor tant windows of exposure or effect lie outside ELS test protocols. Obviously, if there is m eaningful potential for effects outside the ELS exposure period, ELS tests should not be used as surrogates for more involved chronic tests. It may also be that the knowledge of toxicological mechanism for a particular chemical may be sufficient that meaningful chronic toxicity data could be developed from exposures that have a structure different from the life-cycle, partial lif e-cycle, and ELS protocols defined explicitly in

the Guidelines . While defining such alternate protocols is beyond the scope of this document, we recognize the potential for such a situation and leave it to appropriate implementation of the

“Good Science” clause to allow for inclusion of su ch alternative exposure protocols as surrogates for chronic toxicity data, most lik ely in addition to, rather than instead of, data from life-cycle toxicity tests.

At the other end of the spectrum lie toxicity tests that extend beyond the definition of a full life18 cycle test, often referred to as multi-generational tests. Because they encompass the full range of life processes as a life-cycle test, we feel that they should be incl uded as acceptable chronic tests, assuming they meet all other requirements for te st acceptability. Some studies have reported

effects from EDC or other chemicals in which exposure to one generati on creates effects in a later generation that were not observed in prior ge nerations even at the same life stage (Nash et al. 2004). If substantial, such e ffects could create a situation wher e even full life-cycle toxicity tests might underestimate the chronic toxicity of a chemical and therefore produce criteria that are potentially under-protect ive. While we recognize the potential for this s ituation, at present we

believe there is not sufficient reason to make mul ti-generational testing a requirement for criteria development, unless there is specific, compe lling information that a criterion would be substantially under-pro tective if multi-genera tional effects were not rigorously considered.

3.2.4 Selecting Effect Endpoints Upon Which to Base ALC

The selection of endpoints appropriate to the de rivation of ALC must be tied to the narrative

intent of the overall Guidelines . The stated goal of ALC is to “protect aquatic organisms and their uses” (see Water Quality Standards Handbook ; U.S. EPA 1994b). While the exact meaning of “protection” is not defined, there is considerable discussion in the Guidelines document that makes clear that protection does not mean the prevention of any m easurable biological effect in any organism. Instead, there is discussion of endpoints that are “biologically important” and prevention of “unacceptable effects”; this implies th at in the context of cr iteria there are effects that are “biologically unim portant” and/or levels of effect that are “acceptable.” Related concepts include the idea that natural populations can with stand some magnitude/fre quency of disturbance

and still meet the intent of the Guidelines

.

With “protection of aquatic organisms and th eir uses” as the assessment endpoint, a decision must be reached as to which biological respon ses (measurement endpoints in risk assessment parlance) are appropriate to addr ess this goal. Survival, growt h, and reproduction are processes that are generally accepted as being directly rela ted to this goal, as these are all demographic parameters that directly affect population dynami cs (although, the exact qua ntitative relationship is not always fully determined). However, ther e are many more biologica l responses that have been observed in response to toxicant exposure, both at the whole organism level (e.g., behavior) and at lower levels of biological organization (e.g., biochemical or hi stological changes). For many of these endpoints, their relationship to the assessment goal, “p rotection of aquatic organisms and their uses,” is less clear. In this regard, we must consider an additional goal of the

Guidelines – that criteria “provide a reasonable and adequate amount of protection, with only a small possibility of considerable overprote ction or under-protectio n” (page 5 of the Guidelines ). In keeping with this intent, it is important that criteria fo cus on endpoints that affect the assessment endpoint, but not create overprotecti on by preventing any measurable effect (or possibility of that effect). There must be a reasonable, affirmativ e connection between the measured response and the assessment endpoint.

The Agency’s Framework for Ecological Risk Assessment (U.S. EPA 1992) identifies this problem:19 In many cases, measurement endpoints at lower levels of bi ological organization may be more sensitive than those at higher levels. However, because of compensatory mechanisms and other fa ctors, a change in a measurement endpoint at a lower organizational level (e.g., a biochemical alteration) may not necessarily be reflected in changes at a higher level (e.g., population effects). (p. 14)

And later on:

Ideally, the stressor-res ponse evaluation quantifies the relationship between the stressor and the assessment endpoint. When the assessment endpoint can be measured, this analysis is straightfo rward. When it cannot be measured, the relationship between the stressor and meas urement endpoint is established first then additional extrapolations, analyses, and assumptions are used to predict or infer changes in the assessment endpoint. (p. 23)

Measurement endpoints are related to assessment endpoints using the logical structure presented in the conceptual m odel. In some cases, quantitative methods and models are available, but often th e relationship can be described only qualitatively. Because of the lack of standard methods for many of these analyses, professional judgment is an essential compone nt of the evaluati on. It is important to clearly explain the rationale for any analyses and assumptions. (p. 23)

Existing criteria documents contai n many types of data that were not used in the criteria derivation (the documents collate an d review these data, but they ar e not used to actually define the criterion concentration) and it is useful to the discussion here to consider how such data have been interpreted. For example, the following text is derived from the most recent criteria document for ammonia (U.S. EPA 1999, see Appendix 5):

Endpoint indices of abnormalities such as reduced growth, impaired reproduction, reduced survival, and gross anatomical deformities are clinical expressions of altered stru cture and function that origi nate at the ce llular level. Any lesion observed in the test organism is cause for concern and such lesions often provide useful insight into the pot ential adverse clinical and subclinical effects of such toxicants as ammonia. Fo r purposes of protecting human health or welfare these subclinical manifestations o ften serve useful in establishing ‘safe’ exposure conditions for cer tain sensitive individu als within a population.

With fish and other aquatic or ganisms the significance of the adverse effect can be used in the derivation of criteria only after demonstr ation of adverse effects at the population level, such as reduced su rvival, growth, or reproduction. Many of

the data indicate that the concentrations of ammonia that have adverse effects on cells and tissues do not correspondingly cause adverse effects on survival, growth, or reproduction. No data ar e available that quantitatively and systematically link the effects that amm onia is reported to have on fish tissues with effects at the populati on level. This is not to sa y that the investigators who20 reported both tissue effect s and population effects within the same research did

not correlate the observed tissue lesi ons and cellular changes with effects on survival, growth, or reproduction, and a mmonia concentrations. Many did, but they did not attempt to relate their observ ations to ammonia concentrations that would be safe for populations of fish under field conditions nor did they attempt to quantify (e.g., increase in respiratory di ffusion distance associated with gill hyperplasia) the tissue damage and cellu lar changes (Lloyd 1980; Malins 1982). Additionally, for the purpose of deriving ambient water quality criteria, ammoniainduced lesions and cellular changes must be quantified and positively correlated with increasing exposures to ammonia.

In summary, the following have been reported:

1. Fish recover from some histopathologica l effects when placed in water that

does not contain added ammonia.

2. Some histopathological effects are tem porary during continuous exposure of fish to ammonia.

3. Some histopathological effect s have occurred at con centrations of ammonia that did not adversely aff ect survival, growth, or re production during the same

exposures.

Because of the lack of a clear connec tion between histopatho logical effects and effects on populations, histopathological e ndpoints are not used in the derivation of the new criterion, but the possibility of a connection should be the subject of further research.

As discussed in greater detail below, chemical s such as EDCs have been shown to produce a wide variety of measurable changes at many di fferent levels of biol ogical organization. The challenge is to select from among those the endp oints that have sufficien tly clear connection to expected effects on populations of aquatic organisms.

3.2.4.1 Specific Examples of Measurable Changes at Different Levels of Biological Organization

The range of organismal endpoints th at have been reported in the lite rature is vast, and varies to some degree on the organism and toxicant. With resp ect to only the HPG axis in vertebrates, this range of endpoints over and a bove direct measures of surv ival, growth, and reproduction

includes:

• Biochemical measures (e.g., the female-speci fic yolk precursor pr otein vitellogenin; native hormones estradiol, testos terone, 11-ketotestosterone);

• Histopathological measures (e.g., proportion of spermatogonia, pres ence of testis-ova, oocyte atresia, Leydig cell hyperplasia/hypertrophy);21 • Gross morphology (e.g., secondary sex charac teristics: nuptial t ubercles, coloration, ovipositors); and

• Behavioral measures (e.g., nest building, defense/aggression).

A comprehensive survey and eval uation of all such endpoints is far beyond the scope of this document. In lieu of that, this section presents in depth discussion of se veral individual measures relevant to the HPG axis, including their streng ths and weaknesses as direct indicators of likely population level effects. The point of this discussion is simply to provide examples of the issues that must be considered in making a decision as to the biological im portance and scientific defensibility for a specific endpoint, organism, or toxicant as it pertai ns to ALC derivation. These decisions will likely require case by case c onsideration; in certain circumstances, the suitability of a particular endpoint may vary across chemicals depending on how an individual chemical influences that endpoint.

One of the challenges that arises when incorporating alternative endpoints into criteria derivation is the need to not only conclude that the endpoint warrants consideration, but also establish some definition of what level of effect on that endpo int is unacceptable. While these links may not be completely quantitative, one would not want the definition of an unacceptable effect on one endpoint to be grossly di sproportional to that considered unac ceptable for another (i.e., if a 20% reduction in reproduction is considered unacceptabl e, what degree of estradiol (E2) suppression is equivalent to a 20% reduction in reproduction?).

In the text that follows, endpoi nts are categorized as being either “organiz ational” or “activational.” Organizational endpo ints are those that are a result of changes to the normal growth and development of an organism, and ar e generally not reversib le with cessation of exposure. Activational endpoints are those that occur in comparativ ely plastic tissues in response to exposure, but which may revert to their prio r or normal condition with cessation of exposure.

Organizational Endpoint: Sex Reversal

Exposure of developing fish to endocrine-active materials during sensitiv e “windows” in early development can skew phenotypic sex dramatically toward either females (estrogenic chemicals) or males (androgenic chemicals). This response has been exploited by aquaculturists, who for many years have used potent natural or synt hetic steroids to produce mono-sex stocks. The sensitivity of fish to this type of “sex reve rsal” is species-specific, and critical windows of exposure can vary markedly across species. The re sponse can be manifested in several different ways, ranging from more or less completely sex- reversed animals (i.e., occurrence of gonads and secondary sex characteristics completely reflective of the opposite sex) to more subtle changes,

such as the occurrence of inters ex gonads (discussed further be low). A significant challenge in assessing this condition-either in the lab or field-is knowledge of actual genetic sex of the fish.

Since the molecular basis of sex determinati on in many fish is not known, reliable genetic markers of what sex an animal is programmed to be are not available for most test species (one notable exception here is the Japanese medaka, which is commonly used for endocrine testing in some parts of the world; Ankley and Johnson 2004). Th e net result of this is that the only way to practically monitor sex reversal in most fish species is indirectl y, through analysis of sex ratios (generally based on phenotypic sex). This requ ires, of course, considerable background22 knowledge concerning “normal” sex ratios for a species (or even strain) of fi sh. For some lab test species (e.g., fathead minnow), the normal sex ra tio appears to be about 1:1, while for other commonly-tested small fish models (e.g., zebrafish), the ratio can be quite variable (Ankley and Johnson 2004). In field studies, collection of accur ate sex ratio data also can be exceedingly difficult, depending on variables such as sampli ng gear and location and timing of collections.

Changes in the sex ratio of populatio ns of fish, either in the lab or field, can be quite indicative of an endocrine-specific MOA, indi cating exposure to estrogenic or androgenic chemicals (or even

chemicals that block the actions of sex ster oids). Significantly, from a risk assessment perspective, alterations in sex ratio could also have direct implications for spawning success and population viability. The degree to which sex ratios are critical in determining embryo production will vary based on reproductive strategies of the species of concern (e.g., broadcast versus paired spawners); however, from an evol utionary perspective, one would speculate that any departure from normal sex ratios for a speci es/population might be considered maladaptive.

Organizational Endpoint: Intersex

Exposure to certain classes of endocrine-activ e chemicals during critical windows in early development can produce intersex gonads (commonl y termed testis-ova), a situation in which the gonads simultaneously contain both ovarian and testicular tiss ue. Different studies from around

the world have shown an elevated occurrence of intersex fish downstream of municipal effluents containing natural and synthetic steroidal es trogens, including EE2 (WHO 2002). In fact,

collection of intersex fish from the field has been one of the most visible manifestations of the effects of EDCs on fish/wildlife. For example, in a widely publicized USGS study, Blazer et al. (2007) recently reported that in the South Bran ch of the Potomac River and select nearby drainages, more than 80 percent of all the ma le smallmouth bass sampled had oocytes growing

within their testicular tissue. Although histology is required to determine and quantify intersex, the techniques involved are re latively straight-forward. What is more challenging than measurement is interpretation of the results. For example, it appears that some degree of background intersex can occur, even in species held under carefully-contr olled conditions (Grim et al. 2007). The degree of bac kground intersex and sensitivity to chemically-induced intersex appear to be quite species-specific, requiring a thorough understanding of normal gonad differentiation and development in the species of concern.

Even in species for which background intersex is low, there is uncertainty as to the degree to

which the condition could occur and not inte rfere with normal reproductive function. For example, in a field study in the UK, Jobling et al . (1998) noted a wide range of intersex in roach collected, even from the same site, with severi ty of the response rangi ng from occurrence of a few primary oocytes in otherwis e normal testicular tissue to instances where there was a complete absence of sperm ducts in the males. Arguably, the former fish could produce viable sperm, while the latter certainly would not. S o, although intersex is an intuitively reasonable endpoint upon which to base predictions of possible adverse effects of endocrine-active chemicals on reproductive success, determination of the relationship between severity of the condition and production of viable embryos is required to co nduct this analysis.23 Activational Endpoint: Behavior

Although not usually considered a biomarker in the tr aditional sense, behavior is an endpoint that historically has been seldom used for ecologica l risk assessment, including the derivation of ALC. There are several reasons for this: (1) the ty pes of assays used to assess behavior can be quite labor-intensive, (2) many methods for asse ssing toxicant-induced behavior have some degree of subjectivity, (3) many behavioral change s (e.g., gill ventilation in fish) are relatively non-specific in that they do not necessarily reflec t exposure to chemicals with a specific MOA, and (4) translation of behavioral changes into adverse effects on endpoints such as survival, growth and reproduction can be di fficult. Nonetheless, virtually all environmental toxicologists recognize the potential for chemically-induced altera tions in behavior to in fluence the health of individual animals and populations.

There are some compelling reasons to consider behavior as a potentially useful/important endpoint in assessing the ecological risk of certain classes of e ndocrine-active chemicals. First, estrogens and androgens are known to play relatively specific ro les in a variety of reproductive behaviors in fish, including competition for mates, courtship and nest-holding/guarding. Alterations in any of these beha viors theoretically could affect reproductive success and, hence, population status. In recognition of this there have been several recent papers describing straightforward, relatively quantitative assa ys for assessing the effects of endocrine-active substances on fish reproductive behavior. For example, Mar tinovic et al. (2007) condu cted a study in which they showed that male fathead minnows exposed to a relatively low concentration of 17 β

estradiol, and subsequently placed in a competiti ve spawning situation with non-exposed males, failed to compete successfully for nesting sites/ females. Similar types of results have been reported for other fish species e xposed to steroidal estrogens (e.g., EE2; see Part II of this white paper), suggesting that behavioral alterations could be important to consider, especially if they occur at exposure concentrations below those that cause effects on mo re traditional endpoints such as development and egg production.

Activational Endpoint: Sec ondary Sex Characteristics

As described above, exposure of developing an imals to endocrine-active chemicals can alter phenotypic sex, resulting in skewed sex ratios in populations. These organizational changes observed in secondary sex characte ristics (in sexually dimorphic sp ecies) and/or gonads typically are not reversible. However, it al so is possible to alter secondary sex characteristics, usually in a reversible manner, in sexually-mature fish thr ough exposure to endocrine -active substances. For example, estrogens or anti-androgens can reduc e expression of androge n-dependent secondary sex characteristics in males. Similarly, androgeni c chemicals can cause female fish to develop male secondary sex characteristics, such as nup tial tubercles in the fathead minnow or elongated anal fins in the Japanese meda ka (Seki et al. 2006). Alteration s in secondary sex characteristics

are much less useful indicators of endocrine-mediate d responses in test species, such as zebrafish, with limited sexual dimorphism (Seki et al. 2006).

Alterations in secondary sex charact eristics in adult fish can, in some instances, be subtle and somewhat subjective with respec t to interpretation. Fo r example, reductions in the status of24 existing structures in fish (such as nuptial tubercles in male fath ead minnows or anal fin length in male medaka) can be difficult to qu antify. However, when there is a de novo synthesis of

structural characteristics where none previously existed (such as tubercles in female fathead minnows), the response is not only quite specific (i n this case to an andr ogenic MOA), but very easy to detect (i.e., the baseline, control condition is zero).

Although changes in secondary sex characterist ics appear to be reasonable mechanistic biomarkers for some endocrine MOA, their util ity as a predictor of adverse outcomes (e.g., egg production) is uncertain. Specifi cally, given our current under standing of fish reproductive physiology/endocrinology, causative links between secondary sex characteristics and gamete quality would be difficult to defi ne. At best, a correlative associ ation may be identified between the two parameters. For example, in studies with the synthetic androgenic steroid 17 β

trenbolone, egg production appeared to be reduced at about the same test concentration that caused some degree of nuptia l tubercle formation in females (Ankley et al. 2003).

Activational Endpoint: Vitellogenin

Vitellogenin status is probably the most common ly measured endpoint in studies with endocrineactive chemicals in fish. Measurement of the lip oprotein (or its mRNA) is relatively easy via a variety of methods (although most techniques have some degree of species specificity; Wheeler et al. 2005). Production of mRNA ( vtg ) and protein (VTG) in the live r of female oviparous (egglaying) vertebrates is normally stimulated by ac tivation of the estrogen receptor by endogenous estradiol. The protein is released to the plasma and subsequently deposited in the ovary where it forms a key constituent of developing oocytes. VT G levels in male oviparous animals typically are non-detectable due largely to very low circ ulating estradiol concentrations; however, males retain the molecular “machinery” in the liver necessary to produce VTG. Hence, exposure to even relatively low amounts of exogenous estrogen or estrogen mimics can stimulate a marked induction of VTG in males. The response not only is specific and sensi tive (in part due to a baseline of essentially zero), but relatively sustained after exposure, as the males have no mechanism whereby to clear the protein from their blood.

Although vtg and/or VTG induction in male fish is an excellent biomarker of exposure to estrogens (Lattier et al. 2002), th e response appears to have little direct (i.e., causative) value in terms of predicting adverse effects on reproduc tion (e.g., Wheeler et al. 2005). This perhaps should not be surprising given that VTG production in males is not part of any normal physiological pathway. It is possible, however , that correlative relationships between vtg and VTG induction in males by exogenous estrogens (such as EE2) and overall effects on fish population status could be derived (e.g., Kidd et al. 2007). This certainly merits additional study, but at present, it appears that the most technica lly-defensible use of VTG occurrence in male fish is as an indicator of exposure to estrogenic substances in the field and/or confirmation of chemical MOA in laboratory studies.

As opposed to males, VTG has a clear physiological role in females in that the protein is essential to egg production. Con centrations of VTG in female s can be reduced by endocrineactive chemicals that directly or indirectly inhibit steroid (ultimately estradiol) production. For example, aromatase inhibitors such as some conazole fungicides decrease steroid production by25 inhibiting enzymes involved in steroidogenesi s, while other androgenic chemicals like trenbolone decrease steroid production through feedback inhibition in the HPG axis. As a consequence, these classes of endocrine-ac tive chemicals reduce normal VTG production in female fish, thereby reducing fecundity and, ultim ately, affecting population status (Miller et al. 2007). Therefore, in the case of females, VTG status may be effectively used as a biomarker both of exposure and effects. Kidd et al. (2007) f ound that VTG was elevated in female fathead minnows outside of their spawning season. Therefor e, elevated VTG in females outside of the spawning season may also be an important measure of stress.

3.3 Pathways and Receptors Beyond the HPG-Axis

As was explained at the outset, this section (Section 3) has a s ubstantial focus on the HPG-axis not because it is the only MOA that is of concer n in this document, but because it is currently prominent in both social and scientific arenas. Ho wever, it is important to re-emphasize that the

use of HPG-active chemicals as a basis for discus sion does not imply that this is the only group of CECs of concern with respec t to the development of ALC, or the only group for which there may be need for supplementation of the explicit procedures outlined in the Guidelines

.

As an example, the hypothalamic-pituitary-thyr oid (HPT) axis is another endocrine system present in vertebrates that govern s important biological pathways and is potentially subject to disruption. Similar to the role of steroid hormone s in the HPG axis, actions of the HPT axis are mediated through thyroid hormone, which is invo lved in the regulation of metabolic activity, energy consumption and muscular activity in adul t animals, and the regulation of postembryonic or perinatal growth and development in developi ng animals, especially in the central nervous system (Chatterjee and Tata 1992). Thyroid horm one is also responsible for the obligatory

induction and maintenance of me tamorphosis in amphibian and other poikilotherms, and may also play a role in male reproduction (Peterson et al. 1997). Since the actio ns of thyroid hormone are mediated via binding to highly-conserved nu clear thyroid hormone receptors and modulating transcription of specific genes, di sruption of the HPT axis can be disrupted in many ways parallel to those discussed for the HPG axis (Farwe ll and Braverman 2006), and in doing so, create similar challenges for the development of ALC. Only a few of the developmental actions of thyroid hormones, however, are the result of the direct interaction of th e hormone and receptor. Instead, most are indirect via the influence of thyroid hormone on other hormone or growth factors. For example, some of the growth-promoting effects of thyroid hormones on juveniles are indirectly mediated via growth hormone released from the pituitary gland (Chatterjee and Tata 1992).

The amphibian metamorphosis assay is one of the thyroid-relevant in vivo screening assays EPA has developed to detect chemicals that interf ere with the thyroid hormone system. The assay represents a generalized vertebrate model to the extent that it is based on the conserved structure and functions of the thyroid systems, and thus mirrors some of the assays developed and discussed earlier for the HPG axis. This par ticular assay is important because amphibian metamorphosis provides a well-st udied, thyroid-dependent process which responds to substances active within the HPT axis (Fort et al. 2007). The utility of this and other similar HPT-specific assays for development of ALC is predicated on the principle that th e dramatic morphological26 changes that occur during post- embryonic development of vert ebrates are dependent on the normal function of the HPT axis, and that interf erence with this process leads to quantifiable effects (Zoeller and Tan 2007).

Other pathways relevant to this discussion coul d include any of a number of those regulated by different nuclear hormone-type tr anscription factors, such as the progesterone, glucocorticoid and aryl hydrocarbon (Ah) receptors. Of these the Ah receptor is of particul ar interest because it has been well studied and is key to the toxicity of several impor tant environmental contaminants such as dioxins and PCBs. Ah receptor agonists ar e extremely toxic to early life stages of some vertebrates species (e.g., adult fish are at least 10 times less sensitive than early life stages), can induce delayed mortality not captured in short-term (e.g., 96-hour) toxicity tests, and are not very toxic to invertebrates, which lack the recepto r (Cook et al. 1993; Mount et al. 2003; Tanguay et al. 2005). Hence, as is true for HPG-active chem icals, knowledge that a contaminant may be an Ah receptor agonist can help focus testing to determine ecological risk (Cook et al. 1993).

Although the previous systems are generally found in vertebrates bu t not invertebrates, parallel developmental, reproductive, and homeostatic systems exist in i nvertebrates (Lintelmann et al. 2003) and are most likely just as susceptible to disruption by xenobi otic chemicals. In fact, many pesticides are designed explicitly to disrupt bioc hemical pathways specific to invertebrates or sub-groups of invertebrates as a means to reduce effects on non-ta rget (vertebrate) organisms. Some endocrine-mediated processes unique to certa in taxa of invertebra tes include molting, limb generation, diapause, pheromone production, pigmentation and coloration, and metamorphosis. For these processes, the most important endoc rine regulators in arthropods are ecdysone and related compounds (ecdysteroids), which are involved in embryonic development, molting, metomorphosis, reproduction, and pigmentation (Lintelmann et al . 2003). Juvenile hormones in insects and methylfarnesoate in crustaceans ( both belonging to the class of sexual hormones called terpenoids) are also deemed necessary to mediate the regulatory f unctions of ecdysteroids (DeFur et al. 1999). Research on th e effects of CECs on these system s is still in its early stages, but the parallels with other systems that are susc eptible to disruption are clear, and may therefore create similar issues for the development of ALC.27 4.0 SUMMARY AND RECOMMENDATIONS

Through its deliberations, the workgroup conclude d that the basic framework and conceptual underpinnings of the Guidelines apply to CECs as well as other chemicals. Further, the “Good

Science” clause of the Guidelines provides the flexibility to adopt procedures that will produce a technically rigorous and protective criterion. The focus of this report has been the interpretation and adaptation of the principles set forth in the Guidelines with respect to common toxicological characteristics of CECs. In that regard, the wo rkgroup identified a number of possible modifications or alternate interpretations that might aid those de veloping criteria for CECs to do so in a resource efficient manner that takes best advantage of existing knowledge.

Although some of the recommendations involve incr easing flexibility in meeting certain data requirements, the intent is to guide the generatio n of ALC for CECs that have the same technical rigor as 304(a) criteria develope d for other chemicals; these are not methods for “short-cut” criteria. This is a significant point, because an important feature of the Guidelines is defining a minimum technical rigor that crit eria must have; if insufficient information exists to achieve a

minimum level of confidence in the calculated crit erion, then criteria should not be derived. The important consequence of this for risk assessors a nd managers is that when criteria are used to make regulatory decisions, one can have confidence that uncertainty regarding the criterion is not excessive. In other words, criteria derived using the Guidelines are often used as both “walk away values” (i.e., there is high confidence that there is little or no risk when exposures are below criteria) and as indicators of risk (imply ing that effects are likely when criteria are exceeded). If greater uncertainty were allowed in criteria, then the ab ility to use the values in this way would be compromised.

A negative aspect to establishing a minimum level of information for criteria is that there may be chemicals for which regulatory guidance is ne eded, but for which toxicological data are

insufficient to meet the minimum standards of the Guidelines . In such cases, there may still be a need for alternate approaches to derive interim regulatory gui dance values on which to base

decisions that must be made before sufficient information for a complete water quality criterion can be gathered. While much of the discussion in this report might be useful to inform the development of such an approach, it must be em phasized that developing procedures to derive interim regulatory guidance values based on limited toxicity information is a separate matter and would require considerable additional analysis.

The subsequent sections summarize the issues an d recommendations of the workgroup according to

the areas of concern identified above.

4.1 Relevance of Acute Toxicity Effect Levels in Setting ALC for CECs

Some CECs may not be acutely toxi c, or may only be acutely toxi c at environmentally irrelevant concentrations. Thus, if the minimum data require ments for acute toxicity data are not already met by existing data, conducting additional acute te sts might be unwarranted. Indication of lack28 of acute toxicity in key aquatic species might also warrant dir ect calculation of the CCC rather than using the FAV/FACR approach, and thus e liminate the need for th e full suite of acute toxicity tests normally required.

For a CEC of interest, available information shoul d be reviewed to determine if the CMC would be sufficiently higher than the CCC such that developing the CMC is not needed. Exactly how much data is a risk management judgment, a nd probably does not have a unique answer. We recommend that the following information be considered when addressing this issue:

• the amount and phylogenetic spread of acute toxicity data available;

• toxicity data from short-term exposures that may not meet the strict definitions in the

Guidelines of acute toxicity data acceptable for criteria derivation, but from which information on responses to acute exposures can be inferred;

• data on short-term exposures garnered from longer-term exposures;

• information from closely related chemicals thought to have the same MOA that have more robust acute data sets; and

• knowledge of the degree of phylogenetic distri bution of the toxicity pathway of concern.

4.2 Defining Minimum Data Requirements in Terms of Taxonomic Coverage

One consequence of dropping acute testing requirements in criteria derivation is the inability to calculate a CCC using the ACR approach, i.e., as the quotient of the FAV and FACR. In addition, for chemicals with large ACRs, it is like ly that the mechanisms for acute and chronic toxicity differ (Welshons et al. 2003) and that the sensitivity of different taxa to acute and/or chronic toxicity varies widely. Both of these issues introduce uncertainty into the interpretation of ACRs, and probably make it inadvisable to use the FAV/FACR approach. Under such a circumstance, a prudent approach would gene rally be to develop a CCC directly from a sufficiently robust set of chronic data, using the procedures outlined in the Guidelines . If there is insufficient data from actual toxicity tests to fulfill the MDRs to deve lop a CCC directly from chronic toxicity data, a reasonabl e understanding of th e toxicological MOA for the chemical may allow inferences as to what taxa (and endpoints) are most likely to be insensitive, such that measured chronic values for those taxa might not be needed. One important consideration in this process is to avoid an excessive number of taxa es timated to be insensitive, relative to those for

which actual test results are availa ble, and thus to distort the p hylogenetic distribution from that implicit in the MDRs and typical of ALC.

Accordingly, the workgroup recommends that, for chemicals without complete chronic toxicity data sets fulfilling all MDRs, there be an evalua tion of whether sufficient information exists to

conclude that certain taxa w ould not be sensitive to the ch emical. Given the variation in understanding and availability of data likely to ex ist for different CECs, it is presumed that at least initial application of this approach would have to be justified on a chemical-by-chemical basis using appropriate scientific judgment. However, lines of evidence that might be applicable

to this determination include:

• an in depth understanding of the toxicologi cal (or, in the case of drugs, therapeutic)29 MOA;

• information on the basic physiology of ot her taxa in relation to the MOA;

• toxicity data from chronic exposures or othe r relevant experiments that do not meet the strict definitions of acceptable chronic data given in the Guidelines , but from which information on relative taxon sensitivity can be inferred; and

• information from closely related chemicals thought to have the same MOA that have

more robust acute or chronic data sets.

4.3 Use of Non-Resident Speci es in ALC Development

Historically, EPA has not used data derived from toxicity testing with non-resident species in the actual criteria derivation pr ocess. Excluding species simply because they are not resident may be unnecessarily restrictive for the purposes of deriving national crit eria, and may actually increase rather than decrease uncertainty. Because ALC are intended to protect “most of the species, most

of the time” and use distributions of test da ta for point estimation, increasing the species representation in the toxicological database should allow better estimation of species sensitivity

distributions.

The workgroup recommends that some non-resident species be considered for use in criteria derivation calculations, focusing on those specie s with widely used and standardized test

methods and for which there is no reason to be lieve would misrepresent the sensitivity of comparable resident species. Furthermore, we specifically suggest accepting data for zebrafish ( Danio rerio ) and Japanese medaka ( Oryzias latipes ), to reflect international efforts toward data equivalency (Ankley and Johnson 2004). This recommendation pertai ns to the direct use of chronic toxicity data in the calcu lation of a CCC as is currently done for resident species. It is worth noting that even non-resident species that ar e not included in criteria calculations may still provide important information on MOA, sensitivity of endpoints, etc., as expanded on further

below.

4.4 Defining Appropriate Chronic Toxicity Data

The Guidelines state that acceptable chronic tests for criteria derivation are full life-cycle exposures (egg/birth to e gg/birth) for both vertebra tes and invertebrates, as well as partial lifecycle (adult to juvenile) and early life-stage (ELS; egg to juve nile) tests for fish. For chemicals for which sexual development/maturation or re productive effects are of most concern, the

allowance in the Guidelines for using ELS or partial life -cycle fish tests might need

reconsideration. The Guidelines already give priority to life-cycle tests when they are available and show greater sensitivity than other test s. However, other information indicating the importance of sexual development and reproducti on (from other species, similar chemicals, knowledge of the MOA) might also establish a basis for not considering ELS data and for requiring life-cycle or partia l life-cycle tests for fish.

At present, a CCC could be derived for a chemical for which chronic toxicity data for fish are limited to ELS exposures. Because of the importance of sexual maturation and reproduction for30 determining the chronic toxicity of ch emicals like EDCs, the workgroup recommends strengthening the Guidelines such that the chronic toxicity data requirements require that either:

1) Full life-cycle data be availabl e for at least one fish species; or

2) There is a body of experimental informa tion indicating that li fe processes outside the ELS or partial life-cycle expos ure/observation windows would not be important to capturing the important t oxicological effects of the chemical.

We note further that although this report is focused on CECs, this recommendation may be important to implement for all chemicals, not just CECs.

Regarding the latter, we recognize that there ma y be circumstances where the information that shows the sensitivity of different life stages comes from studies that, while scientifically valid,

for some reason do not meet all th e requirements of a valid life-c ycle test as defined in the

Guidelines . Alternatively, there may be data from expe riments that violate other requirements of acceptable toxicity tests under the Guidelines , but still provide insigh t into sensitive exposure periods or life processes. Even though chronic values from such data may not be used directly to calculate a CCC, it seems a reasonable use of such data to evaluate the question of where in the life-cycle there are important windows of expos ure and/or effect, and how that impinges on

criteria derivation.

It may also be that meaningful chronic toxici ty data could be devel oped from exposures that have a structure different from the life-cycle, partial life-cycle, a nd ELS protocols defined explicitly in the Guidelines ; e.g., a short-term (21-day) repr oduction assay with the fathead minnow (U.S. EPA 2001) or a multi-generational study – see example for EE2 reported in Nash et al. (2004). While defining such alternate protocols is beyond th e scope of this document (see Ankley and Johnson 2004 for more detail), we re cognize the potential for such a situation and leave it to appropriate implement ation of the “Good Science” clause to allow for inclusion of such alternative test protocols as surrogates for chronic toxicity data, most likely in addition to, rather than instead of, data fr om life-cycle toxicity tests.

4.5 Selection of Effect(s) Endpoints Upon Which to Base ALC

Although chronic criteria typica lly are based on direct effect s on reproduction, growth, and survival, there may be other endpoints indirectly re lated to these responses that could be useful

for criteria derivation. The selec tion of endpoints appropriate to the derivation of ALC must be

tied to the narrative intent of the overall Guidelines . The stated goal of ALC is to “protect aquatic organisms and their uses.” While the exact mean ing of “protection” is not defined, there is

considerable discussion in the Guidelines document that makes clear that protection does not mean the prevention of any measurable biologica l effect in any organism. Instead, there is discussion of endpoints that are “biologically important” and prevention of “unacceptable effects”; this implies that in the context of cr iteria there are effects that are “biologically unimportant” and/or levels of effect that are “acceptable.”31 Chronic test data and other data should be examined to determ ine whether, for the specific chemical or MOA, endpoints beyond those traditi onally used for criteria derivation may have intrinsic “biological importance” an d therefore could be used as a basis for defining threshold of effect (e.g., sex ratio). Specifica lly, in the context of EDCs:

• Other “endocrine-sensitive endpoints” (e.g., VT G, testis-ova) should be examined to determine whether they can be relied upon as de finitive indicators of other biologically important endpoints (e.g., reproduc tion), with the idea that they may be incorporated into calculation of the criterion. Impor tant sources of this information would include full lifecycle tests in which these other endpoints were measured alongside traditional chronic endpoints, and may include tests with other chemicals with the same MOA (e.g., E2 for EE2).

• If endpoints, such as VTG or test is-ova, are used as direct or indirect indicators of effect, it is critically important that the base line condition (e.g., va riation during normal development) be understood sufficiently to define when changes are biologically meaningful.

• Selection of appropriate endpoi nts (and their associated eff ect thresholds) may, in some instances, transcend “biological importance” (the focus of the Guidelines ) to reflect societal concerns (e.g., physical ap pearance of wild-caught fish).

4.6 Involvement of an Expert Panel

As becomes clear from the preceding issues, deve lopment of appropriate criteria for CECs may be unusually dependent on technical interpretations of a wide range of toxicological information pertinent to specific chemicals. One of the recommendations from a SETAC Pellston workshop (Mount et al. 2003), consistent with much of the a bove, was that expert panels be used to provide professional judgment during stages of the pr oblem formulation and data interpretation associated with criteria development, partic ularly for chemicals with specific MOA. The involvement of the panel would “ensure considera tion of other existing data for the chemical of concern, enable a significant degr ee of up-front technical input, and provide a level of peer

review that should facilitate wider and more ready acceptance of the recommended criteria.” The workgroup agrees with this reco mmendation and suggests that it be incorporated into criteria development of CECs.

To maximize effectiveness, this panel should be convened very early in the criteria development process such that it will be able to assist in problem formulati on, identification of important data, and scoping of particular issues that will be impo rtant. We envision these panels as being formed around specific chemicals, or groups of chemical s with a similar MOA, in order to access the most specialized expertise available.32 5.0 REFERENCES

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## WHITE PAPER

AQUATIC LIFE CRITERIA FOR CONTAMINANTS OF

EMERGING CONCERN

PART II

ILLUSTRATION OF RECOMMENDATIONS USING DATA FOR

17 α -ETHYNYLESTRADIOL (EE2)

Prepared by the OW/ORD Emerging Contaminants Workgroup

June 03, 2008

NOTICE

THIS DOCUMENT IS AN INTERN AL PLANNING DOCUMENT

It has been prepared for the purpose of Research & Development Planning. It has not been formally released by the U.S. Environmental Protection Agency and should not at this stage be construed to represent Agency guidance or policy.2 TABLE OF CONTENTS

1.0 INTRODUCTION ............................................................................................................... ..4

2.0 DESCRIPTION OF THE ACUTE AND CHRONI C TOXICITY DATA SUMMARIZED

FOR EE2........................................................................................................................ ........6

3.0 EE2 DATA EVALUATION A ND CONSIDERATIONS FO R ALC DEVELOPMENT....7

3.1 RELEVANCE OF ACUTE TOXICITY EFFECT LEVELS IN SETTING ALC FOR

EDCS........................................................................................................................... ......7

3.2 USE OF NON-RESI DENT SPECIES IN ALC DEVELOPMENT ..................................8 3.3 MINIMUM DATA REQUIREMENTS RE GARDING TAXONOMIC COVERAGE....9

3.4 DEFINING APPROPRIATE CH RONIC TOXICITY DATA..........................................12 3.5 SELECTION OF EFFECT(S) ENDPOI NTS UPON WHICH TO BASE ALC ...............15

4.0 REFERENCES ................................................................................................................. .....173 TABLE OF TABLES

Table 3.1. Potential GMAVs fo r Application to EE2 CMC. ............................................................7

Table 3.2. Potential Chronic Valu es for Application to EE2 CCC...................................................11 Table A.1. Effects of EE2 on Aquatic Animals (Short-term survival).............................................28 Table A.2. Effects of EE2 on Aquatic Animals (Long-term survival). ............................................29 Table A.3. Effects of EE2 on Aquatic Animals (Growth)................................................................31 Table A.4. Chronic Reproductive Effects of EE2 on Aquatic Animals (Fecundity, Fertility, and Population Growth). .......................................................................................................33 Table A.5. Chronic Reproductive Effects of EE2 on Aquatic Animals (Sex Reversal)...................35 Table A.6. Chronic Reproductive Effects of EE2 on Aquatic Animals (Intersex)...........................37 Table A.7. Chronic Reproductive Effects of EE2 on Aquatic Animals (Sexual Behavior). ............39 Table A.8. Data on Effects of EE2 on Aquatic Animals (Secondary Sexual Characterstics). .........40 Table A.9. Chronic Reproductive Effects of EE2 on Aquatic Animals (Vitellogenin)....................42 Table A.10. Chronic Effects of EE2 on Aqua tic Animals (Other Relevant Endpoints)...................44

LIST OF APPENDICES

APPENDIX A..................................................................................................................... ...............274 1.0 INTRODUCTION

In Part I of this white paper, toxicological ch aracteristics of some contaminants of emerging concern (CECs) important to th e derivation of ambient water qua lity criteria for aquatic life (aquatic life criteria, ALC) were described, and recommendations were made to facilitate ALC

derivation for these chemicals. In Part II of th is white paper, toxicity data for a model CEC, 17 α

ethynylestradiol (EE2), are used to further il lustrate and explore t hose recommendations. Ethynylestradiol was chosen as a model compound for a several reasons. First, it possesses many of the toxicological characteristics described in Part I, and sufficient toxici ty data exist to allow evaluation of the principles underlying the Pa rt I recommendations. Second, toxicological

effects of EE2 have been found both in the la boratory, the source of t oxicological data for criteria development, and in the field, where crite ria are used to enforce the regulatory authorities of the Clean Water Act. Finally, there is intere st in deriving an EE2 ALC, and using EE2 as a basis for discussion should help advance that goal. While acknowledging that interest, it is important to note that the data and discussi on presented are not intended to represent the formulation of an actual ALC, and potential AL C concentrations should not be inferred. The information from the ecotoxicological literature us ed here is for illustrative purposes and should not be considered as comprehens ive, nor have all the data been fully examined for quality and applicability to ALC development.

The synthetic estrogenic steroid EE2 is the active pharmacological component of most oral contraceptives, and acts as a pot ent estrogen receptor agonist in vertebrate s. After use and excretion of the contraceptive, domestic sewage treatment plant (STP) effluents become the primary source of EE2 entering the aquatic envi ronment (Damstra 2002). Kolpin et al. (2002)

found EE2 in 5.7% of 139 streams monitored in the U.S. While the concentrations of EE2 in Kolpin et al. (2002) have been debated (Erics on et al. 2002; Till 2003), other studies have noted concentrations ranging from 0.1 – 5.1 ng/L in su rface waters (as reviewed by Campbell et al. 2006). Overall, it is somewhat uncertain at this time how high environmental concentrations of EE2 may be. Reliable analytical methods for the detection of EE2 have not been in existence very long, nor have they been widely validated in independent multi-laboratory studies. Some modeling efforts by the pharmaceutical industry i ndicate that based on the level of production and use of EE2 in the U.S., concentrations found in effluents should be less than 1 ng/L (Anderson, P.D. and D'Aco, V, personal communication, 2008). Complicating assessment of the possible risk of EE2 is the fact it co-occu rs in STP effluents with the natural steroid hormones estradiol and estrone, though EE2 is ge nerally found at lower concentrations. These three estrogens reportedly account for the majority of the estrogenic activ ity present in domestic wastewater effluents (Desbrow et al. 1998; Snyder et al. 2001), but EE2 is the most potent and resistant to degradation of the th ree (Nash et al. 2004; Gross-Sor okin et al. 2006). Data collected

from fish and surface waters downstream of STPs over the past decade have implicated steroidal estrogens as the primary constituents in domestic effluents leading to the occurrence of intersex fish (Gross-Sorokin et al. 2006).5 The remainder of this part of the white paper co nsists of a brief description of some relevant acute and chronic toxicity data available for EE2 (Section 2) a nd the evaluation of these data with respect to the recommendati ons made in Part I (Section 3).6 2.0 DESCRIPTION OF THE ACUTE AND CHRONIC TOXICITY DATA

SUMMARIZED FOR EE2

Acute and chronic toxicity data we re identified via a literature search and review of relevant articles from EPA’s ECOTOX database in April 2007. This list of potentially useful articles was supplemented with a few additional reports and arti cles as they became published or available. Only those studies with EE2 effect data on in dividual aquatic organisms or their populations were retained. For this particular effort, all endpoints expressing effect s of EE2 at the whole animal and cellular levels were initially consider ed. Because the EE2 data set is relatively large, and because many studies report more than one endpoint of possible consideration for ALC development, the data have been broadly summar ized in an appendix (Appendix A). The tables comprising Appendix A are organized by endpoint, and include separate tables for endpoints typically used to derive ALC (s urvival, growth and reproduction) as well as for other endpoints relevant to the estrogenic mode of action of EE 2. Table A.1 contains the data available on the acute (lethal) toxicity of EE2 to aquatic animals. This table is followed by others containing the chronic (long-term) effects of EE2 on survival (Tab le A.2) and growth of aquatic animals (Table A.3). Tables A.4 through A.9 present data dire ctly (fecundity, fertility) or indirectly (sex reversal, intersex, sexual behavi or, vitellogenin) related to th e effects of EE2 on reproduction. Finally, Table A.10 presents a summary of the significant effects of EE2 on aquatic animals based on other potentially relevant endpoints. In vitro effects were not considered in the data analysis.

Within each table in Appendix A, data are firs t separated by studies where significant effects were observed, and then by studies where signif icant effects were not observed (i.e., where no effect was observed at the highest concentration tested). Each table in the appendix combines data for aquatic vertebrates and invertebrates for both freshwater and saltwat er species, the latter designated by asterisks. All tables are organized by increasing effect concentrations, and all chronic effect endpoints are as reported by the authors.

Many studies of the effects of EE2 on aquatic or ganisms did not use standard toxicity test protocols, particularly those meas uring sublethal responses. This is probably due in part to these studies having been designed for purposes other th an ALC development, such as exploration of toxicity mechanisms, identification of sensitivity windows, bioassay development, etc. Adequate quantification of effect c oncentrations is also difficult for so me of these studies because of the use of widely spaced treatment concentrations and by problems with anal ytical detection of exposure concentrations near the threshold for reproductive effects. While the results from such studies might limit their use in ALC development according to the definitions in the Guidelines

, they were included in this document because they may inform other aspects of criteria derivation, as explained in general terms in Part I and in detail in the sections that follow.7 3.0 EE2 DATA EVALUATION AND CONS IDERATIONS FOR ALC DEVELOPMENT

This section considers th e application of the data in Appendix A toward criteria derivation in the context of the several areas of concern and genera l recommendations identified in Part I of this white paper.

3.1 RELEVANCE OF ACUTE TOXICITY E FFECT LEVELS IN SETTING ALC FOR

EDCS

One of the recommendations from Part I of this document was to determine whether the acute sensitivities of aquatic organisms to a chemical of interest are sufficient, relative to chronic sensitivity and expected exposur es, to warrant derivation of a criterion maximum concentration (CMC) under the Guidelines procedures. This is especially important if there is not sufficient acute toxicity data to meet the minimum data requirements of the Guidelines , in order to avoid wasting resources on unnecessary additional testing. EE2 provides a good example of a chemical having insufficient acute toxici ty data to derive a CMC according to Guidelines procedures, but enough data to demonstrat e that deriving a CMC is not necessary.

Table 3.1 provides information on GMAVs that might be considered in CMC derivation. These

values were derived from Ta ble A.1 for any tests meeting Guidelines requirements, including "greater than" values indicative of the highest tested concentration eliciting less than 50% mortality. For genera without such acceptable te sts, EC50/LC50s from Table A.1 for tests of 24 h duration and from Table A.2 for tests up to 30 da ys were also used. The EC50/LC50 values for these longer tests are designated as "greater than" values to indi cate the expectation that acute EC50/LC50s would be higher. Values for medaka and zebrafish are included in accordance with

the recommendation from Part I of this document th at some latitude be adopted regarding species

not resident to North Americ a. Acute tests with embryos , not usually included in CMC calculations, are also included here because they s uggest greater sensitivity of this life stage.

Table 3.1. Potential GMAVs for Application to EE2 CMC.
Genus GMAV (ng/L) Comments

Freshwater

Rana >760,000 14-d test

Gammarus >840,000 10-d test

Medaka >1,000,000

Danio 1,700,000

Ceriodaphnia 1,800,000

Hydra 3,800,000

Sida >4,100,000 24-h test8 Genus GMAV (ng/L) Comments

Daphnia >5,000,000

Chironomus 9,100,000 24-h test

Saltwater

Lytechinus 30,000 Embryo

Strongylocentrotus 30,000 Embryo

Acartia 88,000 Embryo

Tisbe >100,000 21-d test

Acartia 1,100,000

Neomysis 1,200,000

The data summarized in Table 3.1 show several deficiencies in meeting the minimum data requirements for deriving a CMC under the Guidelin es. For freshwater application, only four genera, rather than the minimum of eight, meet the acute test requirements, even if the prohibition for non-resident specie s is ignored. If the shorter and longer tests are included, the requirement of at least eight genera is met, but the requirement for a salmonid fish is not. Even if this requirement is also ignored, two of the lowest four genera are "greater than" values, whereas CMC calculations can only be made if the four most se nsitive genera have definite values ("greater than" values are permitted only for more tolerant genera.) For saltwater, there are even greater deficiencies in meeting the minimum data requirements.

Although these data are insufficient for deriving CMCs, they do provide ample evidence that a CMC is not needed and that it is unnecessary to conduct further tests to meet the minimum data requirements. For freshwater, there is still a rather broad taxonomi c representation, including three vertebrates from two different classes, four crustaceans from tw o orders, and a third phylum. The acute LC50s/EC50s are consistently near and above 1 mg/L, several orders of magnitude above both the most sensitive chronic endpoints (Tables A.4 – A.9) and the highest

environmental concentrations that organisms might be exposed t o. The saltwater data do show greater sensitivity for the embryonic stages of so me genera, but whether this reflects a lifestage or taxa sensitivity issue, thes e LC50s/EC50s are still four orde rs of magnitude above the most sensitive chronic endpoints and en vironmentally-relevant exposures.

3.2 USE OF NON-RESIDENT SPECIES IN ALC DEVELOPMENT

Under the Guidelines , toxicity values from aquatic species not resident to the contiguous 48 United States, Alaska, or Canada are ex cluded from ALC derivation. One of the recommendations in Part I of this white paper is th at this prohibition be relaxed and that data for non-resident species be allowed where deemed suita ble, especially for species such as medaka

and zebrafish which have become standard te st organisms commonly used worldwide. Any tested species, whether resident or not, serv es as a surrogate for estimating a sensitivity9 distribution relevant to assessing risks in a va riety of aquatic communities with a multitude of untested species. Therefore, the issue here is whether a non-resident species can serve as a reasonable surrogate for assessing the sensitivity of untes ted resident species. The use of such species would still be contraindica ted if there is reason to believ e they are significantly more or less sensitive than resident species.

The data in Appendix A support the use of medaka and zebrafish data in criteria calculations. Although there are no resident fish species with which to compare the ac ute sensitivities of medaka and zebrafish (see Table 3.1 and Table A.1), their lack of acute sensitivity is consistent with that of resident amphibians and invertebrates in the available data. The sensitivities of these fish species for long-term survival (Table A .2), growth (Table A.3) , and reproduction (Table A.4) are interspersed with thos e of resident fish species, so there is no indication of either substantially higher tolerance or sensitivity to contrai ndicate their use. This is also generally true of the other endpoints summarized in Appendix A. The similarity among fish species of different geographic origins is not surprising, since the MOA of EE2 involves receptors and pathways that are highly conserved among vertebrates. If similar trends are seen in the data once they are thoroughly examined for quality and appli cability to ALC development, data from these non-resident species should be included in criteria development.

The data in Appendix A also underscore pragmatic advantages of including non-resident species in criteria development. Medaka and zebrafish provided a large fraction of the available data regarding EE2 effects on fish. Removing them from the dataset simply because they are not

resident would limit information on the distribut ion of species sensitivity and may actually increase rather than decrease uncertainty regardi ng resident species. Another use of data from non-resident species could be to assist in extrap olations of information across species, chemicals, and endpoints. For example, life-cycle tests with medaka could be used to evaluate whether early life-stage or partial life-cycl e tests with resident species s hould or should not be accepted in criteria calculations for specifi c classes of chemicals with a de fined MOA. The relationship of reproductive effects in non-resident fish (Table A.4) to other endpoints (Tables A.5-A.9) could also be used to determine how to apply informa tion on these other endpoints for resident species lacking direct toxicological information on reproduction.

3.3 MINIMUM DATA REQUIREMENTS RE GARDING TAXONOMIC COVERAGE

As discussed in section 3.1, deriving a CMC fo r EE2 is not useful because acutely-toxic

concentrations are so much hi gher than both chronic effects concentrations and expected environmental concentrations. In addition, de veloping a CMC would require additional acute toxicity tests to meet the minimum data requirements (MDRs) specified in the Guidelines

. Without a CMC, the criterion con tinuous concentration (C CC) must be calculated directly from the available data, rather than th rough extrapolation using an acute to chronic ratio (ACR); this is probably not advisable anyway for such larg e ACRs. Since the ACR method is moot, the

Guidelines calculation procedures for the CCC require that there be suffici ent chronic toxicity tests to satisfy the MDRs for estimating the fifth percentile of the chronic database. For10 freshwater criteria, these MDRs include a specie s from: the family Salmonidae; a species from a second family in the class Osteichthyes; a specie s from a third family in the phylum Chordata; a planktonic species from the Class Crustacea; a benthic species from the Class Crustacean; a species from the Class Insecta; a species from a phylum other than Chordata or Arthropoda; and a species from an order of insects or a phylum not othe rwise represented.

Few existing ALC have chronic data that meet the MDRs, and this will likely be true of CECs as well. Significant expense would be incurred co nducting new chronic test s to satisfy all the requirements. As recommended in Part I of th is white paper, because only the four most sensitive genus mean chronic va lues (GMCVs) are used in the criteria calculations, chronic testing requirements for a taxon needed to meet an MDR should be waived if there is sufficient

information to conclude that this taxon is more tolerant than the four most sensitive genera. A value (or values) for this taxon would still be included in the data se t, but its GMCV would simply be specified to be greater than the fourth lowest GMCV.

Table 3.2 lists chronic values for the toxicity of EE2 to various freshwater genera to illustrate data that might be included in freshwater crit eria calculations. These chronic values were obtained from Tables A.2-A.4, using Guidelines data selection procedur es where possible, but

also included some additional data to support discussion of how cer tain data deficiencies might be addressed. For invertebrates, the Guidelines require life-cycle tests that include reproductive endpoints, but if that type of test was not availabl e, then other tests are re ported here, with their limitations noted. For fish, the Guidelines preference order of life-cycl e, partial life-cycle, and early life stage (ELS) was followed, but other tests were also reported as ne eded for illustrative purposes, with their limitations also noted. Fo r all genera, the most sensitive endpoint among chronic survival, growth, and reproduction was se lected, which was from the reproduction data of Table A.4, except for Chironomus (for which development from egg to pre-emergence was

tested and the effects concentration was from Table A.3). Each chronic value (CV) was calculated as the geometric mean of the repor ted no observed and lowest observed effect concentrations (NOEC and LOEC) for an advers e effect. When the LOEC was the lowest exposure concentration, a "less-th an" concentration is reported for the CV and, when the NOEC was the higher exposure concentrat ion for insensitive species, a "g reater-than" concentration is reported. As explained in Sections 1 and 2, th ese data are still under review and subject to modification. The specific values here should not be misconstrued as final, but rather as examples to illustrate trends and indicate needs that support the recommendations being

addressed here.11 Table 3.2. Potential Chronic Values for Application to EE2 CCC. Genus CV(s) (ng/L) Notes

Danio 0.6, 1.5, <1.1 Life-cycle tests; for 1.5 ng/L CV, there was a 9-fold

difference between LOEC and NOEC and the LOEC

was a 100% effect Pimephales <0.32, 1.5 Life-cycle tests; for <0.32 ng/L CV, LOEC showed

reduced fertilization but in creased egg production so total reproduction not adve rsely affected; for 1.5 CV,

4-fold difference between NOEC and LOEC Oryzias 3.2 F 0 from 1 d through spawning; 10-fold difference

between NOEC and LOEC Oncorhynchus <16 Adult exposure only; fertilization success only endpoint examined
Potamopyrgus 50 Adult exposure only; em bryo production over 9 wk

test

Gammarus >7,600 Population size over 100 d test; increased population size at 760 and 7,600 ng/L Daphnia 45,000 5-fold difference between NOEC and LOEC

Tisbe >100,000 Saltwater copepod included to further indicate arthropod insensitivity

Chironomus 320,000 Larval growth and molting schedule only; did not include emergence and reproduction

Brachionus 800,000 Intrinsic rate of population increase over 72 hr test

The data in Table 3.2 indicate high sensitivity of vertebrates to EE2. Significant reproductive effects in the life-cycle tests for zebrafish ( Danio rerio ) and fathead minnow ( Pimephales promelas ) occur at concentrations near and perhaps below 1 ng/L. Although the chronic test for medaka ( Oryzias latipes ) did not cover the entire life cycl e, it included life stages likely important for reproduction and indicated a sensitiv ity similar to zebrafish and fathead minnow. For rainbow trout ( Oncorhynchus mykiss ), a more limited exposure addressing only effects on fertilization suggests that reprodu ctive effects on this species shoul d also be present in the low ng/L range. However, the absence of a definite toxicity value for ra inbow trout will be an important impediment to criteria calculations, bo th for leaving an MDR unsatisfied and for being one of the four most sensitive genera in this se t. Actual criteria development will require a decision whether to (a) require more information for this species, (b) use other information to help estimate rainbow trout sens itivity or (c) justify setting the MDR aside (see Section 3.5).

The invertebrate data in Ta ble 3.2 indicate lower sensitivit y, especially for arthropods ( Gammarus, Daphnia, Tisbe

, Chironomus ) and rotifers ( Brachionus ). Some data, like the

Chironomus test, fail to satisfy the Guidelines requirement for a life-cycle test and the copepod

Tisbe is a saltwater species included here only to reinforce conclusi ons about arthropod sensitivity. Also, the tests for Gammarus and Brachionus are not standard life-cycle tests, but could be considered to satisfy Guidelines requirements because exposures span a life cycle and12 include reproductive eff ects. The snail ( Potamopyrgus ) toxicity test showed moderate sensitivity, although still about an order of magnitude less than the fish, and also does not involve a full life-cycle test.

These data demonstrate the potential for a situ ation in which the GMCV s for taxa reasonably expected to be insensitive do not need to be quantified. For example, although the Chironomus

test was not a full lif e-cycle test and thus could no t fully define the GMCV under Guidelines requirements, it indicates such a degree of insensitivity for growth and development, such that it can be reasonably presumed that a full life-cycle test would still show much less sensitivity than the vertebrates, especially because other arthropods are observed to be similarly insensitive. Likewise, the snail test, although not for a full life- cycle, involved the effects of long exposures on reproduction, and can be argued to be sufficiently less sensitive than fish reproduction so that it would not reasonably be expected to be among the four most sensitive gene ra if a life cycle test

was conducted. These inclusions , along with the data for Daphnia

, Gammarus , and Brachionus

, satisfy the Guidelines MDRs for invertebrates, and would al low an ALC to be calculated from the four sensitive vertebrate genera, provided th e value for the rainbow trout was resolved.

Assessing that taxa are likely to be insensitive could involve othe r lines of evidence, especially for CECs with more limited chronic toxicity da ta than EE2. Tests invo lving endpoints such as

those in Tables A.5-A.10 could be used to es tablish that certain ta xa are sufficiently less sensitive than others to preclude the need fo r tests on their chronic survival, growth, and reproduction (Tables A.2-A.4), the endpoints typica lly used in ALC development. Information from other chemicals might also be used, such as using the insensitivity of arthropods to EE2 to preclude testing this taxonomic class with chemi cals with the same MOA. Such a strategy could be used to help the evaluation of EE2, particular ly regarding the snail se nsitivity. For example, the sensitivity of this or similar species relative to that of vertebrates for other chemicals could be used to strenghten a conclu sion that they are less sensit ive to EE2 than are fish.

3.4 DEFINING APPROPRIATE CHRONIC TOXICITY DATA

As discussed in Part I of this white paper, ch aracteristics of some CE Cs require that careful consideration be given to the selection of chronic toxicity data appropriate for ALC development. Specifically, the use of data from early-life stage (ELS) or partial-life cycle (PLC) exposures as estimates of life-cycle chronic effect thresholds is inadvisable for chemicals whose MOA would result in biological effects for which critical periods of i nduction and/or expression would lie outside the exposure/observati on window provided by the test procedure.

An examination of data specifically for EE2 provides evidence to support emphasis on full lifecycle exposures for determining the chronic toxic ity of EE2. Länge et al. (2001) conducted a full life-cycle chronic exposure with fathead minnow s which were exposed from fertilized eggs (F 0 ) through maturation, spawning, and early-l ife stage development of the F 1 generation. Nominal exposure concentrations of EE2 were 0.2, 1, 4, 16, and 64 ng/L (for convenience, nominal

concentrations are used in this discussion as the importa nt point is relative endpoint sensitivity13 rather than absolute concentrati ons inducing effects). As part of this exposure, measurements of growth (as length) and survival were made at 28 days post hatch (dph) which would correspond to the end of a standard ELS exposure with fa thead minnows. At 28 dph, there were no effects on survival. The length endpoint showed a 16% reduction at 64 ng/L, a smaller but significant reduction of 6% at 16 ng/L, and no effect at 4 ng/L. Accordingly, the NOEC and LOEC for an

ELS test with EE2 would have been 4 and 16 ng/L, respectively, and an EC20 based on length would be >64 ng/L. However, as exposure c ontinued throughout the life cycle, pronounced

effects were observed for other endpoints at lower exposures. There was no reproduction at all in fish exposed to 4 ng/L, and a trend, though not significant, toward lower reproduction at 0.2 and 1 ng/L. Other significant effects observed included a 16% reduction in weight of adult female fish at 1 ng/L after 301 d exposure, a nd 5 to 10 percent reducti ons in weight of F 1 offspring at 28 dph, though the authors questi oned the biological significance of the F 1 growth effects. Regardless, the clear i ndication is that life-cycle exposu re showed substa ntially greater sensitivity to EE2 than was evidenced from ELS endpoints alone. This was much larger than the factor of 2 difference generally found fo r other chemicals by McKim et al. (1978).

A similar conclusion can be drawn from the st udy of Parrott and Blunt (2005). This involved exposure from fertilized egg through reproduction, including measures of fertilization success (but not ELS development) in the F 1 generation. Exposure was to nominal concentrations of 0.32, 0.96, 3.2, 9.6, and 32 ng/L EE2. Measurements of survival and growth at 30 dph showed no effects (NOEC ≥ 32 ng/L). However, continuation of exposure through a dulthood showed no reproduction in the 3.2, 9.6 and 32 ng EE2/L treatmen ts, and all fish in these treatments were phenotypic females. There was also suggestion of effects on fertilization success at 0.32 and 0.96 ng/L, although interpretation of these effects is complicated by an increase in number of eggs produced in these same treatments, such that the total number of fertilized eggs was not as dramatically affected. Regardless, the message rela tive to definition of ch ronic sensitivity is the same in that effects were appa rent after life-cycle exposure at concentrat ions well below those that would be expected to show effects in an ELS test.

Additional comparisons can be extracted from th e work of Wenzel et al. (2001), who conducted a multi-generational study of zebrafish exposed to EE2 concentrations from 0.05 to 10 ng EE2/L. Observations of survival and length of exposed fish showed no effects at 21 and 42 dph (NOEC

≥ 10 ng/L). However, with continued exposure, a variety of effects were observed around an EE2 test concentration of 1 ng/L, including effects on adult length, time to spawning, egg production

and fertilization. As for the fathead minnow st udies, survival and grow th measured during the period comparable to an ELS study were far less sensitive to EE2 exposure than were endpoints measured in full life-cycle st udies (Tables A.2, A.3, A.4).

The reason for these differences between ELS and full life-cycle tests is obvious when one considers the MOA for EE2, wh ich interferes with sexual differentiation, development, maturation, and spawning. Because the endpoints m easured in ELS tests are limited to survival and growth, and because the effect s of EE2 on sexual differentiation are not apparent (at least not at a gross morphological level) in the tested species by 28-30 dph, the ELS test is comparatively insensitive to toxicity mediated through an estrog en receptor signaling pathway. It is interesting14 to note that even though a standard ELS test is relatively insensit ive to detecting the effects of EE2 exposure, other work has shown that key windows of exposure do in fact occur during the ELS exposure window. Van Aerle et al (2002) demonstrated that la rval fathead minnows exposed to EE2 only during brief windows durin g early development (e.g., 10-15 dph) showed altered sexual development of ma le fish at 100 dph, including the development of an ovarian-like cavity and changes in the distri bution of testicular cell types (Table A.6). The issue for interpreting chronic toxicity data is that, even though eff ects may be induced during ELS exposure, they are not expressed unless exposed fish are observed later in sexual development.

This latter observation also has implications fo r the suitability of PLC tests for detecting the effects of EE2 or other chemicals acting thr ough a similar pathway. As discussed in the

Guidelines , PLC tests are acceptable chronic tests for fish species that require more than one year to reach sexual maturity, such as the common speci es of trout. PLC tests are to begin exposure “with immature juveniles at leas t 2 months prior to active gonad developmen t, continue through maturation and reproduction, and end not less than 24 days (90 days for salmonids) after the hatching of the next generation.” If salmonids (or other species for which PLC tests might be conducted) were to express effects from larval exposure to EE2 as observed by Van Aerle et al. (2002) for fathead minnows, one would expect that a PLC exposure would not be as sensitive as a full life-cycle exposure. That is, even though a PLC test includes the observation periods shown to be sensitive in full life-cycle exposures, it might not include the exposure windows important to inducing chronic effects on sexual differentiation and development.

While the rationale for emphasi zing full life-cycle chronic test s has clear grounding in the MOA for EE2, it has practical implications in terms of the fish species likel y to be tested. Most species for which life-cycle chronic tests are most commonly conducted (primarily fathead minnow, zebrafish, medaka, flagfish, and sheepshead minnow), are small fish that develop rapidly and are continuous spawners (as opposed to annual spawners like rainbow trout or bluegill sunfish). Whether or not these life history traits influe nce sensitivity to EE2 is unknown, but because of the investment necessary to conduct true life-cyc le exposures with annually spawning fish that take much longer to develop, it may be unlikely th at comparative data will be developed. Better understanding these implications is a worthy subject for future research.

Finally, there are some scattered indications in th e literature for trans-generational effects of EE2 exposure. As mentioned above, Länge et al. ( 2001) found small effects on growth in the F 1 generation that were not observed after comparable exposure of the F 0 . Wenzel et al. (2002) also report some suggestions of growth inhibition in subsequent gene rations at exposure below those causing such effects in the first exposed generati on (Table A.3). The mechanisms by which such effects might occur are not clear, nor are their im plications (in fact, Länge et al. actively dismiss them as being biologically unimportant). At th is point, it does not seem that the evidence for trans-generational effects is su fficient for requiring their incl usion in the definition of an acceptable chronic test, but the po tential for the existence and im portance of trans-generational

effects should be re-evaluated in the futu re as additional data become available.15 3.5 SELECTION OF EFFECT(S) ENDPOINTS UPON WHICH TO BASE ALC

Aquatic studies with EE2, particul arly those using fish, have meas ured a variety of endpoints not traditionally used for criteria derivation, including reproductive behavior, abnormal sex ratios, changes in secondary sexual char acteristics, altered histopathol ogy (typically gonadal), changes in steroid hormones, and modifications in the expression (or activ ity) of a variety of proteins/enzymes. Many of these endpoints were evaluated because they are known (or hypothesized) to be responsive to estrogenic MO A, and not because the intended result was the quantitative assessment of risk. Among the chal lenges in using data from these types of mechanism-based endpoints is that such measurements are seldom standardized or straightforward in their interpretation. For exam ple, alterations in beha vior are difficult to objectively quantify, it is challengi ng to accurately measure steroid hormone concentrations in small fish, and the capability of measuring gene expression or enzyme activity can be quite lab/method-specific. A second source of uncertain ty in using most of the mechanism-specific endpoints evaluated in EE2 studies is a lack of knowledge con cerning the functional relationship between changes in endpoints a nd responses of primary concern for risk assessment, such as survival, growth and reproduction. Even in cons idering these challenges related to measurement and interpretation, however, there are a handful of mechanistic e ndpoints/responses that exhibit utility for supporting ALC derivation for EE2 (or other xenobiotic compounds with estrogenic activity). Three of these are disc ussed in greater detail below.

A frequently measured mechanism-specific endpo int in fish exposed to EE2, is induction of vitellogenin mRNA ( vtg ) or expression of circulating vitell ogenin protein (VTG) in males (Table A.9). The most attractive attribut e of this endpoint is its specificity for an estrogenic MOA, since there are no other chemically linked biol ogical phenomena known to consistently activate the vitellogenin gene or elevate vi tellogenin protein in male fish. Further, since the vitellogenin gene is quiescent in male fish, which implies a zero baseline of vitellogenin, the response is unambiguous with regard to exposure. Additio nally, this exposure mediated induction of vtg is sensitive to low levels of exogenous estrogen. B ecause vitellogenin protein has been frequently evaluated in fish studies, accura te methods of measurement (including several commercial kits) are available for many fish species, including the small fish models for which much of the EE2 chronic toxicity data exist. Given these attributes, male VTG has and should continue to be a very useful endpoint for monitoring the occurre nce of estrogenic chemi cals (including EE2) in the environment. A major drawback to using ma le-specific circulating pr otein to assess exposure and risk of EE2 (or other estr ogens), including the derivation of ALC, is the lack of an established functional linkage betw een expression of the protein and adverse endpoints related to early development or reproduction (Wh eeler et al. 2005). This is in la rge part due to the fact that VTG plays no physiological role in male reproductive processes. As such, any associations that might exist between VTG induction in males and reproductive success is likely more correlative

than causal.

Despite the fact that th e appearance of VTG in males appear s not to be a r obust predictor of adverse effects on reproduction, the response co uld nonetheless play an important role in reducing the uncertainty of ALC for EE2, or the development of ALC for other chemicals which16 might be estrogenic. From Table 3.2 above, it is appa rent that data from lif e-cycle tests with fish would be appropriate (and critic al) to setting the final ALC for EE2 and, by extension, other estrogens. Hence, knowledge that a less well-studied chemical than EE2 induces VTG in males could be used to help identify those instances wh en one (or more) life-cycle fish assay(s) would be recommended for generating robust data for AL C derivation. Another possible use of protein data, that could have more dir ect applicability to developing an ALC for EE2, involves use of the endpoint as a basis for evaluati ng relative species sensitivity. Specifically, reproductive data suitable for an EE2 ALC are largely from three species: fathead minnow, medaka and zebrafish (Table 3.2); however, there are st udies with numerous species that have evaluated the ability of EE2 to induce vitellogenin mRNA and protein in ma les. Provided that a common dose metric could be established across these studies, a data set could be developed to provide an indication of the relative sensitivity distribut ion of fish species to EE2, in a ddition to other estrogens. This would enable a direct comparison of values along the continuum of estrogen sensitivity for those fish species wherein chronic data exist (based on VTG induction) and, as such, could provide a quantitative indication of uncertain ty for a proposed EE2 criterion.

There are two mechanism-specific endpoints that have been measured in a number of EE2 studies that might, with additi onal research and analysis, have a direct bearing on criteria

derivation: alterations in sex ratio (i.e., generation of genotypic males with a female phenotype)

and the occurrence of intersex/t estis-ova (Tables A .5, A.6). As opposed to VTG induction in males, the functional linkage between skewed sex ratios or abnormal gonad development and reproductive success in fish, at both the individua l and population levels, is readily apparent. Specific endpoints, however, can be difficult to measure. For example, detailed histological analyses are needed to identify a nd, especially, quantify te stis-ova. To detect an alteration in sex ratio, a genotypic marker of gender (available in medaka but not fathead minnow or zebrafish) or a relatively large representative sample is required to reliably de tect chemically-induced changes within a proportion of males and females in a population. Probably more difficult than measurement of the endpoints is definition of the quantitative linkage between changes in sex

ratio or occurrence of testis-ova and eff ects on reproductive success for individuals and populations. For example, unless one assumes th at any deviation in sex ratio from the norm (e.g., 1:1) is adverse, it is nece ssary to know (in the case of estr ogenic effects) the magnitude of shift in respective gender numbe rs that is likely to result in cases where fewer young are produced. Similarly, it is probable that some degree of testis-ova w ould not be considered adverse in terms of reducing re productive success, especially c onsidering that the condition can exist at some degree, even in c ontrol animals (Grim et al. 2007). It is certain that at some level of manifestation, the condition will impair gonad func tion sufficiently such that acceptable levels of normal sperm cannot be produced. The frequenc y of this phenomenon, however, is currently unknown for any fish species. Definition of this relationship would suppor t use of testis-ova occurrence in fish not only for prospective a ssessments (like criteri on derivation), but in environmental monitoring studies focused on chemicals with an estrogenic MOA.

At present, uncertainties regarding measurement and interpretation hamper use of data from any of the mechanism-specific endpoints mentioned a bove as a basis for derivation of an ALC for EE2. Eith appropriate research, however, induction of vitelloge nin in males, changes in sex17 ratios and occurrence of testis-ova, all have the po tential to contribute insi ghts to different facets of quantitative risk assessment for estrogenic chem icals, including derivation of ALC. There is one noteworthy additional observation relative to use of non-traditiona l endpoints for an EE2 ALC. Several fish life-cycle studies using EE2 have been conduc ted in which typical measures of reproductive success (e.g., fecundi ty, fertility) have been made in conjunction with induction of VTG, sex ratio and/or testis-o va data (e.g., Länge et al 2001; Wenzel et al. 2001; Nash et al. 2004; Parrott and Blunt 2005). Although experiment al design variables make some of the endpoint sensitivity comparisons challenging, it do es not appear that there are substantial

differences in EE2 test concentrations that produce adverse effects on egg production/fertility, versus those that alter the mechanism-specific e ndpoints. Hence from a pragmatic perspective, at least for the near-term, it seems reasonable to base an EE2 ALC on traditional measures of long-term reproductive success in fish.

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Effects of EE2 on Aquatic Animals (Short-term survival).Species Life stage Method DurationEC50 or LC50(ng/L) Reference RemarksTraditional Acute (1-7 day timeframe)*Sea urchin,Strongylocentrotus purpuratusembryo S,U 96 h 30,000 Roepke 2005 EC50 - abnormaldevelopment*Sea urchin,Lytechinus anamesusembryo S,U 96 h 30,000 Roepke 2005 EC50 - abnormaldevelopment*Copepod,Acartia tonsaegg R,U 5 d 88,000 Andersenet al. 2001 EC50-Inhibition ofnaupliar developmentMedaka,Oryzias latipesadult - 96 h >1,000,000 Thompson 2000*Copepod,Acartia tonsa10-12 d adult S,U 48 h 1,100,000 Andersen et al. 2001*Opossum shrimp,Neomysis integerJuv, 2-4 mm R, U 96 h 1,200,000 Verslycke et al. 2004Zebrafish,Danio rerioadult - 96 h 1,700,000 Wenzel et al. 2001Cladoceran,Ceriodaphnia reticulataS,U 24 h 1,800,000 Jaser et al. 2003 EC50 mobilityCnidarian,Hydra vulgarisAdult male R, U 96 h 3,800,000 Pascoe et al. 2002Cladoceran,Sida crystallinaS,U 24 h >4,100,000 Jaser et al. 2003 EC50 mobilityCladoceran,Daphnia magna<24 h S,U 48 h >5,000,000 Goto and Hiromi 2003Midge,Chironomus riparius4thinstar S,M 24 h 9,100,000 Lee and Choi 2006*Indicates saltwater species.29Table A.2. Effects of EE2 on Aquatic Animals (Long-term survival).Species Life Stage Method DurationNOECsurvival(ng/L)LOEC -survival(ng/L) Reference RemarksSignificant Effect ObservedZebrafish,Danio rerio1 d old R,U 38 d 10 100 Orn et al 2006 100% mortality at LOECZebrafish,Danio rerio2 dph R,U 58 d 10 100 Hill and Janz 200390% mortality at LOEC (45%control mortality); excess solventMedaka,Oryzias latipes1 d R,M LC: 85 to 110 dph 29 290 Metcalfeet al. 2001 83% mortality at LOEC*Sheepshead minnow,Cyprinodon variegatusjuv F,M PLC:59-7 dph F1 120 330 Zillioux et al. 2001 50% mortality at LOEC (42 days)Medaka,Oryzias latipes6 mo. F,M 21 d 260 490 Seki et al. 200242% mortality at LOEC (4 of 5dead males)Rainbow trout,Oncorhynchus mykiss1+ year F,M 62 d pre-spawning 130 750 Schultzet al. 2003 100% mortality at LOEC (57 days)Medaka,Oryzias latipes4 mo. R,U 14 d 500 2,000 Thompson 2000; Tiltonet al. 2005Significant mortality at LOECZebrafish,Danio reriofert. eggs R,U 5 wk - 5,000 Ortiz-Zarragoitia et al.200650% mortalty of exposed animals*Copepod,Tisbe battagliai<24 h R,U 10 d - >100,000 Hutchinson et al. 1999 Value is an LC50Wood frog,Rana sylvaticaGosner 26 R,U 14 d - 560,000 Hogan et al. 2006 Value is an LC50Amphipod,Gammarus pulex3-5 mm R,M 10 d - 840,000 Watts et al. 2001 Value is an LC50Leopard frog,Rana pipiensGosner 26 R,U 14 d - 890,000 Hogan et al. 2006 Value is an LC50Leopard frog,Rana pipiensGosner 36 R,U 14 d - 1,200,000 Hogan et al. 2006 Value is an LC50No Signficiant Effects Observed (NOECEquals Highest Test Concentration)*Sand goby,Pomatoschistus minutusjuv F,U 7 mo 6 - Robinson et al. 2003Fathead minnow,Pimephales promelasfert eggs F,U 125 d 10 - Parrot et al. 2003Fathead minnow,Pimephales promelasjuvenile F,M 21 d 20 - Panter etal. 2002 Conc. only 40% of nominal30Species Life Stage Method DurationNOECsurvival(ng/L)LOEC -survival(ng/L) Reference RemarksChinese rare minnow,Gobiocypris rarus7 mo F,U 28 d 25 - Zha et al. 2007 10% mortalityZebrafish,Danio reriofert. eggs F,U 3 mo 25 - Van den Belt etal. 2003 40% mortality after 5 mo. recoveryFathead minnow,Pimephales promelasmature male F,U 35 d 50 - Schmid et al. 2002Snail,Potamopyrgus antipodarumadult R,U 9 wk 100 - Jobling et al. 2004Medaka,Oryzias latipes1 d R,U 2 mo 100 - Scholz and Gutzeit 2000 8% mortalityZebrafish,Danio rerio4 wk R,U 33 d 100 - Versonnen and Janssen20046.6% mortality; excessive carriersolventGuppy,Poecilia reticulate(male) < 7 d F,M 108 d 110 - Nielsen and Baatrup 2006Sturgeon,Acipenser fulvescens1 yr F,M 25 d 120 - Palace et al. 2001Rainbow trout male,Oncorhynchus mykissF,M 3 wk 130 - Hook et al. 2007African clawed frog,Xenopus laevisadult R,U 4 wk 2,960 - Urbatzka et al. 2007Wood frog,Rana sylvaticaGosner stage 25 R,M 76 d 4,100 - Mackenzie et al. 2003AmphipodGammarus pulexmixed age F,M 100 d 7,600 - Watts et al. 2002Amphipod adult,Hyalella aztecapre-copulatory R,M 10 wk - 2 xgen 10,000 - Vandenbergh et al. 2003*Copepod,Tisbe battagliai<24 hr R,U 21 d 100,000 - Hutchinson et al. 1999*Copepod,Tisbe battagliai<24 hr R,U 21 d 100,000 - Pounds et al. 2002Cladoceran,Sida crystallinaR,U 34 d 500,000 - Jaser et al. 2003Cladoceran,Daphnia magnaR,U 25 d 500,000 - Goto and Hiromi 2003*Indicates saltwater species.31Table A.3. Effects of EE2on Aquatic Animals (Growth).Species Life Stage Method DurationNOECgrowth(ng/L)LOECgrowth(ng/L) Reference RemarksSignificant Effects ObservedZebrafish ,Danio reriofert egg F,M LC – F1 0.10 0.3 Wenzelet al. 2001 7% reduction at LOEC (75 dph)Zebrafish ,Danio reriofert egg F,M PLC 0.30 1.1 Wenzel etal. 2001 2% reduction at LOEC (78 dph)Zebrafish,Danio rerio20 dph R,M 40 d 0.60 1.5 Orn et al. 2003Increasedjuvenile wet weightFathead minnow,Pimephales promelas<24 hr F,M LC 0.76 2.8 Länge et al. 2001Zebrafish ,Danio reriofert egg F,U 3 mo 1 10 Van den Belt et al. 2003Zebrafish,Danio rerio2 dph R,U 58 d 1 10 Lin and Janz 2006Fathead minnow,Pimephales promelasembryo F,M 114 d - 12 Bogers et al. 2006bChinese rare minnow,Gobiocypris rarus7 mo F,U 28 days 5 25 Zha et al. 2007Fathead minnow,Pimephales promelasfert eggs F,U 60 dph 10 32 Parrot and Wood 2002Zebrafish ,Danio rerio4 wk R,U 33 d 10 100 Versonnen and Janssen 2004 Excessive carrier solventGuppy,Poecilia reticulata<7d F,M 108 d 44 112 Nielsen and Baatrup 2006Increasedadult wet weight.Medaka,Oryzias latipes1 day old R,M LC 29 290 Metcalfe et al. 2001Midge,Chironomus riparius4th instar S,U 48 hr 50 500 Lee et al. 2006 Increasedlarval dry weightMidge,Chironomus riparius1st instar R,M egg - pupa 100,000 1,000,000 Watts et al. 2003No Signficiant Effects Observed (NOECEquals Highest Test Concentration)Zebrafish,Danio reriofert egg F,M 2xGen 4.5 - Nash et al. 2004*Three-spined stickleback,Gasterosteus aculeatusfry R,U 14 days 7.3 - Hahlbeck et al. 2004b32Species Life Stage Method DurationNOECgrowth(ng/L)LOECgrowth(ng/L) Reference RemarksFathead minnow,Pimephales promelasfert eggs F,U 60 dph 10 - Parrot et al. 2003Fathead minnow,Pimephales promelasjuv F,M 21 days 20 - Panter et al. 2002Prosobranch mollusc,Potamopyrgus antipodarumadult R,U 9 wk 100 - Jobling et al. 2004*Sheepshead minnow ,Cyprinodon variegatusjuv F,M PLC 330 - Zillioux et al. 2001Wood frog ,Rana sylvaticaGosner stage 25 R,M 76 d 4,100 - Mackenzie et al. 2003*Indicates saltwater species.33Table A.4. Chronic Reproductive Effects of EE2 on Aquatic Animals (Fecundity, Fertility, and Population Growth).Species Life Stage Method Duration EndpointNOEC(ng/L)LOEC(ng/L) Reference RemarksSignificant Effects ObservedFathead minnow,Pimephales promelas40-60 h F,M LC Percent fertilized ofeggs laid - 0.32 Parrott and Blunt 2005Zebrafish,Danio reriofert egg F,M LC Number of fertilized eggs per female 0.30 1.1 Wenzel et al. 2001Fathead minnow,Pimephales promelas<24 hr F,M LC Mean no. eggs laid per breeding day 0.76 2.8 Länge et al. 2001Zebrafish,Danio reriofert egg F,M 118 d No. eggs spawned and prop fertilized - 3 Fenske et al. 2005Fathead minnows,Pimephales promelas- Field 3 yrs Population crash - 3.2 – 8.9 Kidd et al. 2007Green frog,Rana clamitansfert egg Field 2 yr Hatching success - 3.2 – 8.9 Park and Kidd 2005Zebrafish,Danio reriofert egg F,M 2 x gen Proportion of non-viable eggs 0.50 4.5 Nash et al.2004 Complete Rep. failure at LOEC*Sand goby,Pomatoschistus minutusjuv F,U 7 months Fertile eggs and hatching success - 6 Robinson et al. 2003Fathead minnow,Pimephales promelas6-11 mo. F,M 3 wk Fert rate and no. eggs spawned 0.75 7.5 Pawlowski et al. 2004IncreasedNo. eggs spawned upto 0.75 ng/LZebrafish,Danio rerio2 dph R,U 60 d % viable eggs, % hatch, % swim-up 1 10 Hill and Janz2003 Excessive carrier solventMedaka,Oryzias latipes1 d R,U 2 mo Female egg production 1 10 Scholz and Gutzeit 2000 No effect on male fert at 10 ng/LZebrafish,Danio reriofert egg F,U 3 mo No. spawning females & egg prod 1 10 Van den Belt et al. 2003Zebrafish,Danio rerio8 mo R,U 14 d Absence of intact eggsin ovaries 1 10 Versonnen and Janssen2004Rainbow trout,Oncorhynchus mykiss1+ year F,M PLC Fertilization success - 16 Schultz et al. 2003 EC50; same response@131 ng/LSnail,Potamopyrgus antipodarumadult R,U 9 wk Embryo production 25 100 Jobling et al. 2004 EE2 at 25 ng/L stimulatory*Sheepshead minnow,Cyprinodonvariegatusjuv F,M PLC Hatching success 18 120 Zillioux et al. 2001Medaka,Oryzias latipes6 mo. F,M 21 d Fecundity 260 490 Seki et al. 200234Species Life Stage Method Duration EndpointNOEC(ng/L)LOEC(ng/L)Reference RemarksMedaka,Oryzias latipes4 mo. R,U 14 d Spawning frequency, % fertilizedand % hatch.5 500 Thompson 2000; Tilton etal. 2005Cladoceran,Daphnia magna- R,U 25 d Embryo production 20,000 100,000 Goto and Hiromi 2003Rotifer,Brachionus calyciflorusS,U 72 hr Ratio of ovigerous/non-ovigerousfemales202,000 510,000 Radix et al. 2002Rotifer,Brachionus calyciflorusS,U 72 hr Intrinsic rate population increase r 510,000 1,300,000 Radix et al. 2002No Signficiant Effects Observed (NOEC Equals Highest Test Concentration)Fathead minnow,Pimephales promelas>6 mo. F,M 3 wk No. spawnings and eggs per spawn 1.5 - Brian et al. 2007Mink frog,Rana septentrionalisfert egg Field 2 yr Hatching success 3.2 – 8.9 - Park and Kidd 2005Zebrafish,Danio rerio(females) 5-6 mo R,U 15 d Sterility infemales 5,000 - Ortiz-Zarragoitia et al.2006Amphipod,Gammarus pulexmixed ages F, M 100 d Population growth(total pop. size) 7,600 - Watts et al. 2002Increasein population size*Copepod,Tisbe battagliai<24 hr R,U 21 days Fecundity 100,000 - Hutchinson et al. 1999*Copepod,Tisbe battagliai<24 hr R,U 21 days Reproduction 100,000 - Pounds et al. 2002*Indicates saltwater species.35Table A.5. Chronic Reproductive Effects ofEE2 on Aquatic Animals (Sex Reversal).Species Life Stage Method Duration EffectNOEC(ng/L)LOEC(ng/L) Reference RemarksSignificant Effects ObservedZebrafish,Danio reriofert egg F,U 3 mo Delayed sexual differentiation - 0.10 Van den Belt et al. 2003 No males at LOECZebrafish,Danio rerio20 dph R,M 40 d Male:female sex ratios - 0.60 Ornet al. 2003 Complete sex reversal at 1.5 ng/LFathead minnow,Pimephales promelas40-60 h old F,M LC Male:femal sex ratio 0.32 0.96Parrott and Blunt 2005 Complete ex. femin. at 3.5 ng/LFathead minnow,Pimephales promelasfert. eggs F,U 60 d Male:female sex ratio 0.32 1.0Parrot and Wood 2002 Complete ex. femin. at 3.2 ng/LZebrafish,Danio rerio2 dph R,U 58 d Male:female sex ratio - 1.0 Lin and Janz 2006Fathead minnow,Pimephales promelas<24 hr F,M LC Sex reversal - all female 0.76 2.8 Länge et al. 2001 Sex ratio at 0.76 ng/L 54:46Zebrafish,Danio reriofert egg F,M 42 d Male feminization - 3.0 Fenske et al. 2005Fathead minnow,Pimephales promelasfert eggs F,U 125 d Sex reversal- all female - 10 Parrot et al. 2003Zebrafish,Danio rerio1 dph R,U 60 d Complete feminization - 10 Orn et al 2006Zebrafish,Danio rerio1 dph R,U 60 d Complete feminization - 10 Yamani 2004Fathead minnow,Pimephales promelasembryo F,M 114 d 75% female gonads; 15% undeveloped- 12 Bogers et al. 2006bZebrafish,Danio reriofert egg R,M 60 d Complete feminization - 15 Andersen et al. 2003bMedaka,Oryzias latipes1 d R,M LC Male:female sex ratio 2.9 29 Metcalfe et al. 2001*Three-spined stickleback,GasterosteusaculeatusLarvae R,U 42 d Sex reversal and intersex - 50 Hahlbeck et al. 2004aMedaka,Oryzias latipes1 d R,U 2 mo Sex reversal withovary 10 100 Scholz and Gutzeit 2000Medaka,Oryzias latipes1 d R,U 60 d 88% female, 2% male, 10%intersex10 100 Yamani 2004 and Orn et al2006Amphipod,Gammarus pulexmixed ages F, M 100 d Male:femal ratio- 104 Watts et al. 2002 No dose-response >104 ng/L36Species Life Stage Method Duration EffectNOEC(ng/L)LOEC(ng/L)Reference RemarksGuppy,Poecilia reticulate< 7 d F,M 108 d Male:female sex ratio 44 110 Nielsen and Baatrup 2006No Signficiant Effects Observed (NOEC Equals Highest Test Concentration)Green frog,Rana clamitansfert egg Field 2 yr Male:female sex ratio 3.2 – 8.9 - Park and Kidd 2005Mink frog,Rana septentrionalisfert egg Field 2 yr Male:female sex ratio 3.2 – 8.9 - Park and Kidd 2005Amphipod,Hyalella aztecaadult R,M 10 wk; 2 gen Male:femalsex ratio 10,000 - Vandenbergh et al. 2003*Copepod,Tisbe battagliai<24 hr R,U 21 d Male:female sex ratio 100,000 - Hutchinson et al. 1999Cladoceran,Daphnia magna- R,U 25 d Male:female ratio 500,000 - Goto and Hiromi 2003*Indicates saltwater species.37Table A.6. Chronic Reproductive Effects of EE2 on Aquatic Animals (Intersex).Species Life Stage Method Duration EffectNOEC(ng/L)LOEC(ng/L) Reference CommentsSignificant Effects ObservedPearl dace,Margariscus margaritamature Field 3 yr Presence of testis-ova- 3.2 – 8.9 Palace et al. 2006 Edema in ovariesFathead minnows,Pimephales promelasField 3 yr Presence of testis-ova - 3.2– 8.9 Kidd et al. 2007 Testicular malformationsMink frog,Rana septentrionalisfert eggs Field 2 yr Intersex gonads(5 – 12 %) - 3.2 – 8.9 Park and Kidd 2005Chinese rare minnow,Gobiocypris rarus7 mo F,U 28 d Testis-ova in males 1.05.0 Zha et al. 2007 No sperm detectableFathead minnow,Pimephales promelasegg F,U 5 d Ovarian cavities in males (8%) - 10 Van Aerle et al. 2002Fathead minnow,Pimephales promelas5-10 dph F,U 5 d Ovarian cavities in males (38%) - 10 Van Aerle et al. 2002Fathead minnow,Pimephales promelas10-15 dph F,U 5 d Ovarian cavities in males (64%) - 10 Van Aerle et al. 2002Fathead minnow,Pimephales promelas15-20 dph F,U 5 d Ovarian cavities in males (43%) - 10 Van Aerle et al. 2002Fathead minnow,Pimephales promelasegg F,U 20 d Ovarian cavities in males (22%) - 10 Van Aerle et al. 2002Medaka,Oryzias latipes1 d R,M LC Sex inversion and testis-ova 2.929 Metcalfe et al. 2001 4 of 4 males with TO*Three-spined stickleback,GasterosteusaculeatusLarvae R,U 42 d Intersexed gonads - 50 Hahlbeck et al. 2004aMedaka,Oryzias latipes6 mo F,M 21 d Testis-ova in males (33%) 33 64Seki et al. 2002 No histological abnormalities infemalesMedaka,Oryzias latipes1 d R,U 2 mo All males developed an ovary 10 100Scholz and Gutzeit 2000 No effect on male fertilityMedaka,Oryzias latipes1 d R,U 60 d Intersexed gonads (10%) 10 100 Yamani 2004 and Orn etal 2006Guppy,Poecilia reticulata< 7 d F,M 108 d Feminization of malereproductive ducts44 110 Nielsen and Baatrup 2006Amphipod,Hyalella aztecafert eggs R,M 2 x gen Oocyte-like structures in males 23 320 Vandenbergh et al. 2003Leopard frog,Rana pipiensGosner 25 R,M 162 d Intersex and altered testiculardevelepment414 4,140 Mackenzie et al. 200338Species Life Stage Method Duration EffectNOEC(ng/L)LOEC(ng/L)Reference CommentsWood frog,Rana sylvaticaGosner 25-28R,M 76 d Intersex and altered testiculardevelopment- 4,140 Mackenzie et al. 2003No Signficiant Effects Observed (NOEC Equals Highest Test Concentration)Green frog,Rana clamitansfert eggs Field 2 yr Intersexedgonads 3.2 – 8.9 - Park and Kidd 2005Zebrafish,Danio rerio2 dph R,U 58 d Testis-ova 10 - Lin and Janz 2006Zebrafish,Danio rerio2 dph R,U 58 d Testis-ova 10 - Hill and Janz 2003 Excessive carrier solvent; highcontrol mortalityZebrafish,Danio rerioadult F,U 21 d Feminization oftestes 25 - Islinger et al. 2003*Indicates saltwater species.39Table A.7. Chronic Reproductive Effects ofEE2 on Aquatic Animals (Sexual Behavior).Species Life Stage Method Duration EffectNOEC(ng/L)LOEC(ng/L) Reference RemarksSignificant Effects Observed*Sand goby,Pomatoschistus minutusjuv F,U 7 mo Male nesting behavior - 6 Robinson etal. 2003 High mortality in solvent controls in firstmonth of exposureFathead minnow,Pimephales promelasmature F,M 27 d Impaired ability to compete andacquire territory2.0 8.9 Majewski et al. 2002*Three-spined stickleback,Gasterosterusaculeatusmature R,U 12 d Time spent near nest and glueingfrequency, but effect short-lived- 10 Brian et al. 2006No Signficiant Effects Observed (NOEC Equals Highest Test Concentration)Zebrafish,Danio reriofert egg F,M 2 x gen Natural spawning behavior of adultmale fish4.5 - Nash et al. 2004 Sexually compromised males still activelyparticipated in the spawning act, i.e.,chasing females and competing withhealthy malesAmphipod,Gammarus pulex3-5 mm R,M 10 d Pre-copulatory guarding behavior 3,700,000 - Watts et al. 2001Reproductive behavior was only disruptedat high concentrations where it would beunrealistic to attribute effects to andendocrine-mediated process.40Table A.8. Data on Effects of EE2 on Aquatic Animals (Secondary Sexual Characterstics).Species Life Stage Method Duration Effect Event AssociationNOEC(ng/L)LOEC(ng/L) Reference CommentsSignificant Effects ObservedFathead minnow,Pimephalespromelas6-11 mo. F,M 3 wk Number of male nuptialtuberclesActivational - 0.80 Pawlowski et al. 2004Fathead minnow,Pimephalespromelas40-60 h F,M LC Ovipositor size, nuptialtubercles, banding strengthOrganizational 0.32 0.96 Parrott and Blunt 2005Fathead minnow,Pimephalespromelas156 dph F,M LC Nuptial tubercles, bandingstrength, dorsal fin dot, dorsalfat padActivational 0.32 0.96 Parrott and Blunt 2005Fathead minnow,Pimephalespromelasfert eggs F,U 60 dph Male sex index: nuptialtubercles, dorsal fat pad, dorsalfin dot, banding strengthOrganizational 0.32 1.0 Parrot and Wood 2002 Complete femin. at 3.2 ng/LFathead minnow),Pimephalespromelas<24 hr F,M LC Secondary sex characteristics –not specifiedOrganizational 0.76 2.8 Länge etal. 2001 50% sex ratio at 0.76 ng/LFathead minnow,Pimephalespromelasfert eggs F,U 60 dph Development and length ofovipositorsOrganizational 1.0 3.2 Parrot and Wood 2002 Complete femin. at 3.5 ng/LZebrafish,Danio reriofert egg F,M 2 x gen Coloration and bright anal finmarkingsOrganizational - 4.5 Nash et al. 2004*Sand goby,Pomatoschistus minutusjuv F,U 7 mo Delayed and inhibited nuptialcoloration in malesActivational - 6.0 Robinson et al. 2003Fathead minnow,Pimephalespromelasfert eggs F,U 125 d Ovipositor size, nuptialtubercles, banding strengthOrganizational - 10 Parrot et al. 2003Fathead minnow,Pimephalespromelasmaturing F,M 21 d Number and prominence ofnuptial tubercles and dorsal fatpadActivational - 11 Filby et al. 2007Fathead minnow,Pimephalespromelasembryo F,M 114 d Number and prominence ofnuptial tuberclesOrganizational - 12 Bogers et al. 2006bAmphipod,Hyalella aztecafert eggs R,M 2 x gen Male secondgnathopods Organizational - 23 Vandenberghet al. 2003 No effect >1,000 ng/LNo Signficiant Effects Observed (NOEC Equals Highest Test Concentration)Fathead minnow,Pimephalespromelas>6 mo. F,M 3 wk Relative fat pad weight,number of nuptial tubercles,nuptial tubercle prominenceActivational 1.5 - Brian et al. 2007Medaka,Oryzias latipes4 mo R,U 14 d Anal and dorsal fin shape Organizational 500 - Thompson 2000; Tiltonet al. 200541*Indicates saltwater species.42Table A.9. Chronic Reproductive Effectsof EE2 on Aquatic Animals (Vitellogenin).Species Life Stage Method Duration EffectNOEC(ng/L)LOEC(ng/L) Reference CommentsSignificant Effects ObservedZebrafish,Danio rerioadult males F,M 40 d Increased whole-blood VTG level - 0.50 Nash et al. 2004Fathead minnow,Pimephales promelas6-11 mo F,M 3 wk Increased plasma VTG levels - 0.80 Pawlowski et al. 2004Rainbow trout,Oncorhynchus mykissimmaturefemaleF,M 14 d Increased livervtgmRNA expressionand plasma VTG0.21 1.0 Thomas-Jones et al. 2003Zebrafish,Danio rerio20 dph R,M 40 d Increased whole body VTG levels - 1.5 Orn et al. 2003Zebrafish ,Danio rerioadult male R,U 21 d Increased plasma VTG level - 1.6 Fenske et al. 2001Rainbow trout,Oncorhynchus mykissAdult male F,M 3 wk Increased plasma VTG level - 1.8 Jobling et al. 1996Zebrafish,Danio reriofert egg F,M 42 d Increased vitellogenin level - 3.0 Fenske et al. 2005Fathead minnow,Pimephales promelas- Field 5 mo Increased plasma VTG levels - 3.2 – 8.9 Palace et al. 2002 (alsosee Kidd et al. 2007)Pearl dace,Margariscus margarita- Field 3 yrs Increased whole body VTG level - 3.2 – 8.9 Palace et al. 2006Zebrafish,Danio rerioadult male F,M 8 d Increased whole-body VTG level 2.2 3.6 Rose et al. 2002 EC10 = 0.92 ng/LZebrafish,Danio rerioadult females F,M 40 d Increased whole-blood VTG level 0.50 4.5 Nash et al. 2004Fathead minnow,Pimephales promelasjuv F,M 21 d Increased whole-body VTG level 2.0 5.0 Panter et al. 2002Fathead minnow,Pimephales promelas8 mo R,M 48 hrs Increased livervtglevels 2.5 5.0 Biales et al. 2007Ide,Leucisus idusjuv F,M 7 d Increased plasma VTG levels - 6.0 Allner et al. 1999Zebrafish,Danio reriofert egg R,M 4 d Increased whole bodyVTG level 2.6 7.8 Bogers et al. 2006aZebrafish,Danio rerioadult females R,M 21 d Increased plasma VTG level 4.1 8.5 Van den Belt et al. 2004Zebrafish,Danio rerioadult males F,M 24 d Increased plasmaVTG level - 9.0 Van den Belt et al. 200243Rainbow trout,Oncorhynchus mykiss11 mo F,M 2 wk Increased plasm VTG level 0.87 10 Samuelsson et al. 2006*Eelpout,Zoarces viviparusadult female F,M 3 wk Increased plasma VTG level 5.0 10 Korsgaard et al. 2002Fathead minnow),Pimephales promelas<24 hr F,M LC Increased whole body VTG level 2.8 12 Länge et al. 2001Zebrafish,Danio rerioadults F,M 168 hrs Increased plasma VTG level - 14 Hoffmann et al. 2006Sturgeon,Acipenser fulvescens1 yr F,M 25 d Increased plasma VTG levels - 14 Palace et al. 2001Lake trout,Salvelinus namaycushimmature F,M 21 d Increased plasma VTG level - 15 Werner et al. 2003 Excessive carrier solvent*Baltic flounder,Platichthys flesusadult F,M 21 d Increased plasma VTG level in maleand female fish- 15 Allen et al. 1999bMedaka,Oryzias latipes6 mo. F,M 21 d Increased liver Vtg levels 33 64 Seki et al. 2002Rainbow trout,Oncorhynchus mykissjuvenile R,M 14 d Increased plasma VTG level 10 100 Verslycke et al. 2002*Sheepshead minnow,Cyprinodonvariegatusmale F,M 16 d Increased livervtgmRNA expression 24 110 Folmar et al. 2000Rainbow trout,Oncorhynchus mykiss mature male F,M 61 d Increased plasma VTG levels - 140 Skillman et al. 2006No Signficiant Effects Observed (NOEC Equals Highest Test Concentration)Zebrafish,Danio rerioadult males F,M 310 dpfF1No effect on whole-blood VTG level 4.5 - Nash et al. 2004Zebrafish,Danio rerioadult females F,M 310 dpfF1No effect on whole-blood VTG level 4.5 - Nash et al. 2004*Indicates saltwater species.44Table A.10. Chronic Effects of EE2 on Aquatic Animals (Other Relevant Endpoints).Species Life Stage Method Duration Effect Concentration (ng/L) Reference RemarksSignificant Effects ObservedZebrafish ,Danio rerio17-20 dpf R,U 3 d Enhancedeffect on CYP19A2 geneexpression0.30 Kazeto et al. 2004 Excessive carrier solventZebrafish,Danio rerio2 dph R,U 58 d Suppression of gametogenesis for males(no testes discernable) and females1.0 Weber et al 2003 Excessive carrier solventChinese rare minnow,Gobiocyprisrarus7 mo F,U 28 d Increased GSI and renal somatic index(RSI) in males1.0 Zha et al. 2007Fathead minnow,Pimephales promelas40-60 h F,M LC Reduced GSI in females 3.5 Parrott and Blunt 2005Atlantic salmon,Salmo salarimmature S,U 7 d Increased AchE and GST activities andlactate content after 3 days, but no effectat 7 days5.0 Greco et al. 2007Chinese rare minnow,Gobiocyprisrarus7 mo F,U 28 d Reduced GSI in females 5.0 Zha et al. 2007Chinese rare minnow,Gobiocyprisrarus7 mo F,U 28 d Increased RSI in females 5.0 Zha et al. 2007Medaka,Oryzias latipes4 mo R,U 14 d Increased male and female plasma E2levels5.0 Thomposn 2000; Tilton et al.2005Fathead minnow,Pimephales promelas6-11 mo F,M 3 wk Reduction on maleGSI 7.5 Pawlowski et al. 2004Zebrafish,Danio rerioAdult female R,M 21 d Reduced female ovariansomatic index 8.5 Van den Belt et al. 2004Rainbow trout,Oncorhynchus mykiss11 mo F,M 2 wk Higher hepatosomaticindex (HSI) 10 Samuelsson et al. 2006Fathead minnow,Pimephales promelasfert eggs F,U 125 d Increased liver somatic index 10 Parrot et al. 2003Medaka,Oryzias latipes1 d R,U 4 mo In both males and females, significantlyincreased number of necrotic hepatocytesand kidney tubule cells10 Weber et al. 2004*Baltic flounder,Platichthys flesusadult F,M 21 d Increased HSI in males 15 Allen et al. 1999Zebrafish,Danio rerioadult male S,M 21 d Increased levels of cyp19a2 mRNA(aromatase)21 Kallivretaki et al. 2006Chinese rare minnow,Gobiocyprisrarus7 mo F,U 28 d Increased HSI in males 25 Zha et al. 200745Species Life Stage Method Duration Effect Concentration (ng/L) Reference RemarksChinese rare minnow,Gobiocyprisrarus7 mo F,U 28 d Ovary degeneration in females 25 Zha et al. 2007Zebrafish, Danio rerio adult male F,M 7 d Decreased testrosterone and 11-ketotestosterone levels26 Andersen et al. 2006Atlantic salmon,Salmo salarimmature S,U 72 hr Induced expression of brain P450Aromatase50 Lyssimachou et al. 2006Sturgeon,Acipenser fulvescens1 yr F,M 25 d Increased plasma Vit E, A1 and A2;Decreased Vit E and A in kidney60 Palace et al. 2001*Sheepshead minnow,Cyprinodonvariegatusjuv F,M PLC Increased pathological condition ofkidneys120 Zillioux et al. 2001 Fish survived toreproductionRainbow trout,Oncorhynchus mykissmature male F,M 3 wk Changed geneexpression profile 130 Hook et al. 2007Zebrafish , Danio rerio 18-21 d R,U 72 hr Stimulated expression of CytochromeP450 aromatase (Aro-B)300 Le Page et al. 2006Medaka,Oryzias latipesmature R,U 14 d Induced ER protein and aromatase activity 500 Contractor et al. 2004Medaka,Oryzias latipesadult R,U 14 d Increased hepatic estrogen receptor (ER) 500 Thompson 2000Medaka,Oryzias latipes4 mo. R,U 14 d Decreased female and male GSI 500 Thompson 2000; Tilton et al.2005African clawed frog,Xenopus laevisadult R,U 4 wk Reduced Leutinizing hormone B mRNAexpression3,000 Urbatzka et al. 2006African clawed frog, Xenopus laevis adult R,U 4 wk Reduced testosterone levels in both sexes 3,000 Urbatzka et al. 2007African clawed frog, Xenopus laevis adult R,U 4 wk Reduced E2 level in females3,000 Urbatzka et al. 2007Midge,Chironomus riparius4th instar S,U 24 h Increased expression of heat shockproteins8,000 Lee et al. 2006Medaka,Oryzias latipesmature male R,U 6 d Increased mRNA expression of liverchoriogenin L10,000 Lee et al. 2002bMedaka,Oryzias latipesmature male R,U 6 d Increased mRNA expression of liverchoriogenin H levels20,000 Lee et al. 2002bMedaka,Oryzias latipesmature male R,U 6 d Increased mRNA expression of liverchoriogenin H levels20,000 Lee et al. 2002aMedaka,Oryzias latipesjuv R,U 6 d Increased mRNA expression of wholebody Choriogenic H50,000 Lee et al. 2002aNo Signficiant Effects Observed (NOEC Equals Highest Test Concentration)46Species Life Stage Method Duration Effect Concentration (ng/L) Reference Remarks*Sand goby,Pomatoschistus minutusjuv F,U 7 mo No effect on GSI in males or females 6.0 Robinson et al. 2003Rainbow trout,Oncorhynchus mykissjuvenile R,M 14 d No effect on GSI or HSI 100 Verslycke et al. 2002African clawed frog,Xenopus laevisadult R,U 4 wk No effect on Gonadotropin ReleasingHormone mRNA expression3,000 Urbatzka et al. 2006*Indicates saltwater species.