letter transmitting comments regarding the preliminary ... · 1. page 1: paragraph 2 and table 4....

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RECEIVED March 24,1995

James Colbert Pi a a £7 1995

SPPD BRAXCB K1CI0M VII

Superfund Branch USEPA Region VII 726 Minnesota AveKansas City, Kansas 66101

Re: Comments on EPA's Preliminary Identification of Contaminants of Concern.Ecological Risk Assessment for Mississippi River Pool 15 and Procedures for the Ecological Risk Assessment at the Mississippi River Pool 15 Site . Davenport. Iowa

Dear Mr. Colbert:

Enclosed are three copies of our comments regarding the above-referenced documents. These comments represent a compilation of efforts by Alcoa, Woodward- Clyde Consultants (WCC) and Geraghty & Miller (G&M). In general, we feel the approach presented in the referenced procedures document is only a screening-level assessment and is inconsistent with EPA's Framework for Ecological Risk Assessment and current risk assessment methodology. It is unclear from the procedures document how the screening assessment is intended to be used and there was a notable absence of procedures for "risk characterization". Significant components of the approach are missing, such as the process of problem formulation.

EPA is allowing Alcoa to conduct the ecological risk assessment in the outfalls (USEPA 1994a), along with human health and ecological risk assessments related to other facility units. The procedures document appears to also include the outfalls as part of EPA's Mississippi River Pool 15 (MRP 15) risk assessment which introduces significant redundancy. We do not believe the risk assessments for MRP 15 and the wetlands can logically be conducted independently of the outfalls and other on-site units since these areas involve the same sources, potential contaminants of concern and pathways of contamination. Perhaps more important, we do not feel it is appropriate to separate MRP15, the wetlands, outfalls, and other on-site Facility Site Assessment (FSA) Units near the river from an ecological standpoint since they are contiguous habitats.

We request that Alcoa be permitted to perform the human health and ecological risk assessments of MRP 15 and the wetlands as a natural extension of the work already being conducted on the outfalls and other on-site FSA Units. This will enable the outfalls, wetlands and MRP 15 to be evaluated concurrently in the context of appropriate habitats and contaminants of concern. This will also assure a consistent approach which will be difficult or impossible to accomplish if separate assessments are conducted. f A ^Site; P & l

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To assure a quality and consistent approach to the risk assessments, I would have a joint team from Alcoa, G&M and WCC work on the assessments. Key risk- assessment members of this team would include Dr. Frank Jones and Linda Lawton from G&M, Dr. John Conner, Dr. Douglas Reagan, Carl Crane and Richard Young from WCC; and Kirk Gribben from the Alcoa Remediation Work Group. Enclosed are resumes for key individuals from G&M and WCC. We feel the current partnership between EPA and Alcoa in regards to on-site unit risk assessment activities forms a solid foundation and good example for taking a similar approach in regards to the MRP 15 and wetland assessments.

The comprehensive risk assessments discussed above would be conducted by Alcoa and its consultants with oversight and concurrence of approach from EPA. In fact, we suggest that the process begin with a meeting to discuss the general approach and proposed plans. We would also like to propose that Dr. David Charters, USEPA Environmental Response Team, Edison, NJ, be invited to become involved in the EPA review process at Davenport. Dr. Charters is one of the authors of EPA's draft document “Ecological Risk Assessment Guidance for Superfund: Process for Designing and Conducting Ecological Risk Assessments” (USEPA 1994b). Dr. Charters is currently involved in ecological risk assessment activities at Alcoa's Pt. Comfort, Texas, facility (EPA Region VI) and has expressed to Alcoa’s Kirk Gribben a willingness to participate at Davenport.

We appreciate the opportunity to review the above-referenced document. We would like to have a meeting to discuss these comments in more detail, either in Kansas City or Davenport, once you have had an opportunity to review them.

Yours truly,

Marshall Sonksen, P.E. Location Remediation Manager

cc: Alice Waldhauer - w\ end.Kirk Gribben - w\ end.Jim Bollenbacher - w\o end. Carl Crane - w\o end.

T ■ » •

COMMENTS ON:Preliminary Identification of Contaminants of Concern, Ecological Risk Assessment,

Mississippi River Pool 15 RI/FS Oversight Riverdale, Iowa (Prepared by Jacobs Engineering Group, Inc. for USEPA)

1. Page 1: Paragraph 2 and Table 4. Please provide a reference for “typical”background levels of inorganic chemicals.

2. Page 1: Fourth bullet under Criteria Unavailable. Please define “low toxicity”. If these factors are to be used for defining contaminants of potential ecological concern (COPECs) at other Alcoa units, the criteria need to be more clearly defined to prevent confusion and arbitrary decision making.

3. Page 1: Fourth bullet under Criteria Unavailable. Please define “questionable toxicity”. See comment 2 above.

4. Table 4: The table and text should indicate why chromium VI standards are used as opposed to chromium III. If chromium VI values were used as a conservative measure, this should be stated.

5. Tables 4 and 5: The National Oceanic and Atmospheric Administration (NOAA) Effects Range Low (ERL) and Effects Range Medium (ERM) values are presented for determination of COPECs in sediments. The source of these values referenced is the 1991 NOAA paper. The ERL and ERM values were recalculated by Long et al. (1993) after adding additional, more recent data and refining the approach. The revised NOAA screening values (NOAA 1994) should be used for the determination of COPECs. Some of the values are higher than those presented in Tables 4 and 5; others are lower. The use of the more recent NOAA data will influence the selection of COPECs.

6. Table 3 lists volatile COPECs in surface water based on comparison to available criteria or LOELs. No criteria or LOELs are listed for many of the compounds (e.g., bromodichloromethane, 1,2-dichloroethane; tetrachloroethene; trichloroethene; or vinyl chloride). The NOAA quick reference cards (NOAA 1994) provide values for each of these compounds. These should be used in identifying COECs.

7. Table 6: Many VOCs were identified as COPECs in sediments based simply on their detection in at least one sample. VOCs in sediments should not be construed as a concern in sediments unless there is a continuing source (this will be confirmed in future sampling efforts).

Comments. EA\Mai95 1

COMMENTS ON:

Procedures for the Ecological Risk Assessment at the Mississippi River Pool 15 Site Davenport, Iowa (Prepared by Jacobs Engineering Group, September, 1994, for USEPA)

GENERAL COMMENTS

Alcoa is currently conducting ecological risk assessment of the outfalls as individual site units. The Procedures for the Ecological Risk Assessment at the Mississippi River Pool 15 Site Davenport, Iowa (the Procedures) document appears to include the outfalls as part of EPA’s risk assessment. However, EPA has allowed Alcoa to conduct ecological risk assessment in the outfalls as individual site units (USEPA 1994a). As a result, this introduces significant redundancy. Assessment of MRP15 cannot logically be conducted independently of the outfalls and wetlands since they represent sources and pathways of contamination. Perhaps more importantly, it is not possible to separate the areas from an ecological standpoint since they are contiguous habitats. With oversight from EPA, we propose that Alcoa conduct the ecological risk assessment of MRP15 to appropriately address sources, pathways and habitats in an integrated fashion and to maintain consistency between site and MRP 15 assessment approaches.

OBJECTIVES

There is no clear statement of objectives. The introduction indicates this to be a "quantitative baseline ecological risk assessment" to be used as a "screening tool to provide an initial measure of exposure to site-related contaminants". Page 2 states that "The purpose of this baseline ecological risk assessment is not to calculate the potential risk for each species at MRP 15, but to provide a broad evaluation of the potential risk for several species by calculating the exposure pathways of a few representative species. ...Once the calculations have been completed, the risk characterization section will discuss how the calculated potential risk of an individual ROC may reflect the potential risk of the functional group which it represents." The document acknowledges in Section 5.3.1, that "it is extremely important that the comparison of ADDs and endpoints not be interpreted as a quantification of risk" but rather as "providing an initial understanding of the potential sources of risk". Section 5.3.2 continues on the same note: “As with the terrestrial species, the comparison of endpoints and concentrations of COECs should not be interpreted as a quantification of risk to ROCs. This comparison can only reasonably be interpreted as providing a basic understanding of the potential sources of risk.” Thus, the document in its current form, is a screening-level assessment only. The term quantitative may be more appropriately applied to the "risk characterization", which is alluded to, but not described in the Procedures document.

A clear statement of objectives should be provided. For example, the objectives of the proposed screening assessment may be to focus a more quantitative risk characterization by:(1) refining the list of contaminants of ecological concern; and (2) identifying data gaps and uncertainty that may be addressed through field sampling and/or laboratory testing. In

Comments. EA\Mar95 2

addition, the "risk characterization" portion of the assessment, which is alluded to but not presented, should be clearly described in the current document, or stated that it will be presented as a separate document. Objectives of the quantitative risk characterization might be to establish priorities and provide a scientific basis for remedial decision making with respect to ecological concerns.

Data Quality Objectives

There is no discussion of data quality objectives (DQOs). The DQO process is a systematic planning approach to "ensure that the type, quantity, and quality of environmental data used in decision making are appropriate for the intended application" (USEPA 1993). The DQO process should be incorporated into the ecological risk assessment by identifying the types of decisions that will be made and the associated data requirements, as well as focusing the assessment based on problem formulation (see the following section on "Approach").

APPROACH

There is no reference made in the Procedures document with respect to EPA’s ecological risk assessment guidance. The Framework for Ecological Risk Assessment (USEPA 1992 [the Framework]) should be used as a basis for the overall strategy, especially regarding the elements associated with problem formulation. The Framework document represents EPA’s current approach to ecological risk assessment. Since the analysis and risk characterization steps are based on problem formulation, the results and conclusions drawn from this risk assessment would be impacted.

Problem formulation begins once the constituents of ecological concern (COECs) are determined. Problem formulation includes 5 steps (EPA 1992): Stressor Characteristics [corresponding to Section 2.2.1 of EPA 1992]; Ecosystems Potentially at Risk [2.2.2]; Ecological Effects [2.2.3]; Endpoint Selection [2.3] and Conceptual Models [2.4], Although parts of these sections were included within the Procedures document, overall problem formulation is incomplete because other portions of these sections were either missing or are incomplete. For example:

(1) Stressor Characteristics: No attempt was made to identify all potential physical stressors or chemical stressors that are not site related, or for those chemicals that are site related, to identify other potential sources of these chemicals. This information is important for the completion of the conceptual site models and should be known prior to initiation of the analysis or risk characterization steps.

(2) Ecosystems Potentially at Risk: The ecosystem is broadly referenced asMississippi River Pool 15 (MRP15) and the outfalls and wetlands. Ecosystems potentially at risk should be more distinctly defined in terms of different habitats, transport and exposure pathways, and the spatial distribution of contamination.


(3) Ecological Effects: The Procedures document proposes to use surrogateorganisms (from an ecological standpoint surrogates are guild species or a representative species for a group of organisms that occupy similar ecological niches) to assess impacts to populations of related species. This approach represents an organism-level exposure assessment. However, no explanation was given as to how the risk assessors would extrapolate from organism-level effects, based on a literature review, to risk to populations, communities or ecosystems at the site. The ecosystem, community, population and individual should be defined spatially and temporally. For example, all of MRP15 may be defined as the population for aquatic receptors; this may be different for terrestrial receptors. How will temporal and spatial characteristics of contamination be considered in the quantification of potential ecological risk?

The assessment method proposed in this document is based on a comparison of an applied daily dose (ADD) to a no-observed-adverse-effect level (NOAEL). We agree with the use of NOAELs in a screening level approach for refining COCs and identifying data gaps, but NOAELs should not be used in risk characterization. An exceedance of a NOAEL may still result in no effect to the ecological communities of concern (as will be discussed in the following discussion entitled Endpoint Selection). Some of the questions that arise include:

What is the impact to the ecosystem, aquatic community or to fish populations at the site if a fish NOAEL is exceeded?

How well will organism-level effects depict risk to populations, communities or to the ecosystem in general?

Will field methods be employed to verify an impact if the results of the literature search indicate potential ecological risk?

The questions posed above should be addressed in the Procedures document. Methods that will be used for extrapolating to and interpreting population, community and ecosystem effects should be clearly presented. The document should also outline the decision criteria that will be used to determine whether or not the need exists for additional information (i.e., continued investigation) should the screening assessment indicate potential risks.

(4) Endpoint Selection: Endpoint selection is not adequate to assess ecological risk. Assessment and measurement endpoints were not explicitly identified or discussed. Assessment endpoints were implied only by the identification of functional groups and two types of toxicological impacts: survival and reproduction. These functional groups were assumed to represent the critical ecological components of the ecosystem at risk at the site. These types of assessment endpoints are too broad-based to give specific direction for development of testable hypotheses, which is the basis for the


analysis and risk characterization steps. Measurement endpoints were limited to the identification of no-observed-adverse-effect levels (NOAELs) for a dose or concentration that would be associated with the assessment endpoints. NOAELs are not measurements endpoints but are ecotoxicological values. Consequently, NOAELs may be good for establishing regulations or standards, or may be used for screening

1 level purposes, however they are poor predictors of ecological effects because they do not identify when there is a risk, only when there is no risk. NOAELs cannot be used to identify the magnitude of an effect. This is important considering that population, community or ecosystem level effects are determined based on a “magnitude” of risk and are used for interpretation of ecological significance. Endpoints, in particular assessment endpoints, should be more specific. For example, the assessment endpoint should identify the level of organization that is of concern (ecosystem, community, population, or individual) and the effect of concern. Each assessment endpoint should include within its development an evaluation of ecological relevance. Suter (1993) identified 5 criteria that should be met for endpoint selection: 1) societal relevance; 2) biological relevance; 3) unambiguous operational definition; 4) accessibility to prediction and measurement; and 5) susceptibility to the hazardous agent. Such an analysis was not performed in the determination of endpoints for the site, consequently, the relevance and interpretation of the results of the analysis, when conducted, will be questionable. Proper endpoint selection will allow for development of testable hypotheses that are more focused on the pathways and receptors of greatest concern at the site.

The assessment of ecological risk should be based on a weight-of-evidence approach and not simply on an assessment of toxicity (or a lack of toxicity using NOAELs) derived from a literature search. LOAELs expressed as a percentage of an effect are better indicators of risk to individual organisms, and make extrapolation of effects to populations, communities or ecosystems easier than do NOAELs.

(5) Conceptual Site Models: The document does not present any conceptual site models. Conceptual site models are working hypotheses regarding the release of chemicals from source media, transport of chemicals from source media to exposure media and the biological mechanisms associated with exposure. The identification of conceptual site models associated with different chemical groups is critical to the understanding of the types of receptors and exposure pathways of potential concern for that chemical or group of similar chemicals. For example, the conceptual model associated with volatile compounds released to an aquatic system would be very different than that of PCBs since the primary fate of volatiles released to aquatic systems is volatilization. The fate of PCBs would be sedimentation, bioaccumulation and possibly biomagnification (a preliminary conceptual site model for PCBs was presented by Alcoa in the Sediment/Soil Investigation Studies Work Plan [YMA 1991]). The specific chemicals within these groups also behave differently and these differences should be taken into account in the development of conceptual site models. The lack of conceptual site models does not allow identification of release


and transport mechanisms, elements of exposure and completed exposure pathways for the site — information that is critical for formulation of testable hypothesis (the basis for the analysis and risk characterization steps).

Ecotoxicological Benchmarks

Where NOAELs were not available, safety factors were applied. The use of cumulative safety factors greater than 100 in the derivation of ecotoxicological benchmarks should be an indication that insufficient information is available. The Procedures should state that 100 is the maximum cumulative safety factor. We also believe (as do Suter [1992] and Opresko et al. [1994]) that it is inappropriate to extrapolate (using safety factors) across functional groups as was done in the Procedures. This leads to meaningless values that do not even provide a qualitative indication of risk.

Selection of Receptors of Concern

A preliminary list of receptors of concern (ROCs) was presented in Preliminary Identification of Receptors of Concern Ecological Risk Assessment for Mississippi River Pool 15. The selection of ROCs should be revised as the overall process for the ecological risk assessment evolves. Receptors should be ecologically relevant with respect to habitat availability and importance in ecosystem structure (the ecological relevance of the American toad may be confined to the importance to other toads, rather than to the ecosystem as a whole). Ecological relevance should be assessed based on an ecological inventory of MRP 15. As COECs are selected, the ROCs also should be toxicologically relevant (in the case of PCBs, higher trophic levels are important due to bioaccumulation potential). The process of developing conceptual site models will aid in highlighting receptors that may be of particular concern. Societal relevance may also play a role in the selection of certain ROCs (for example, the bald eagle).

We suggest that the list of ROCs be reviewed, evaluated and redefined, if necessary, during the evolution of the ecological risk assessment process.

Documentation

The Procedures document is to serve as the basis for the ecological assessment of MRP 15. The document will also serve as the framework for ecological assessments at other units of concern. To that end, the exposure assumptions, ecotoxicological benchmarks, and surrogate organisms chosen and presented in the document should be thoroughly referenced, and well documented. Review of the document indicates that this was not done. Several equations and assumptions used in the exposure assessment were not referenced or fully explained so the reader was not able to check the accuracy or validity of the assumptions. Documentation of ecotoxicological benchmark selection was inconsistent and incomplete. Secondary references were often used for ecotoxicological benchmark and criteria development rather than the original document. The criteria development procedures used in these secondary


references (e.g., Suter et al., 1992) were not consistent with the safety factor scheme presented in the document. For consistency, these derived criteria should not be used in the document without further explanation or documentation. Though likely unavailable during preparation of the Procedures, a series of documents that were referenced (for example in Tables 4 and 5) have been superseded by more recently published information (e.g., Mabrey and Suter 1994; Opresko et al. 1994). Data used in the Procedures should be updated as

necessary.

APPLICATION

The document goes through the procedures for calculating screening level risks for a few representative species. It does not indicate how the results of the screening risk assessment will be used, or ultimately how potential risks to the ecosystem at MRP15 will be characterized. We assume that potential risk will be quantified in what the document refers to as the risk characterization. Decision-making criteria used in the screening assessment should be stated. For example:

• if the applied daily dose (ADD) exceeds the ecotoxicological benchmark, then the constituent would be carried into a quantitative risk characterization or an effects assessment; if not, then the constituent would no longer be considered a contaminant of concern.

• if the ADD exceeds the ecotoxicological benchmark, then site-specific sampling may be warranted in cases where a high level of uncertainty was inherent due to assumptions or data gaps.

All assumptions and benchmarks should be re-evaluated in the "risk characterization" with more realistic inputs, as is indicated for food exposures in Section 4.1 ("In the riskcharacterization....... a more reasonable maximum exposure via a particular food source willbe estimated for those pathways which indicate a potential risk"). Each of these points should be clearly discussed in the Procedures document.

SUMMARY

In summary, the Procedures document presents a screening-level assessment. The document does not adequately follow EPA’s Framework for Ecological Risk Assessment because many of the elements that are critical for formulating testable hypotheses (the basis for the quantitative risk characterization) are not included or are poorly defined. Because the Procedures document is not consistent with the Framework, the evaluation and interpretation of ecological significance at the site will be impeded (the final step in risk characterization). Interpretation of ecological significance is critical for making risk management decisions for the site. Exposure assumptions, equations and ecotoxicological benchmarks presented in the document need to be more fully documented and referenced. The steps required to interpret the results of the "screening" assessment, and how the results of "screening" assessment will

Comments. EA\Msr95 7

be used, should be fully outlined. Since this information will affect further investigation and remediation decisions at the site, the "risk characterization" process should be fully incorporated into the Procedures document and agreed upon prior to commencing with the risk assessment.


SPECIFIC COMMENTS

1. Page 1, Section 2.0: "The six ALCOA outfalls and the two shoreline wetlands in the area provide habitats for a wide variety of species." EPA has allowed Alcoa to conduct ecological risk assessment in the outfalls as individual site units (USEPA 1994a). The Procedures document appears to include the outfalls as part of EPA’s risk assessment which introduces significant redundancy. The ecological risk assessment for MRP 15 cannot logically be conducted independently of the outfalls and wetlands since they represent sources and pathways of contamination, and it is not possible to separate the areas from an ecological standpoint since they are contiguous habitats. As indicated in our general comments, we propose that Alcoa conduct the ecological risk assessment of MRP 15 with oversight from EPA to appropriately address sources, pathways and habitats in an integrated fashion and to maintain consistency between site and MRP15 assessment approaches.

2. Page 3: The first paragraph of Subsection 4.1 (Terrestrial Environment - Ingestion Exposure) starts by mentioning that "Many species have a diverse diet." In fact, among higher vertebrates in particular, there are virtually no truly "specialized" feeders in the sense of continuous or sustained use of a single specific type of food (and the few valid examples are not among the ROCs identified for this study). However, for the sake of conservatism, the approach will assume that each ROC essentially eats only one prey type. It is unclear how this will be used in interpretation of potential risk. For example, if one assumes that a mink consumes only mussels in it’s diet, and the resulting applied daily dose (ADD) exceeds a toxicity reference value, will this provide any information other than the relative risk of mussels as a food source? If this has some utility in assessing ecological risk, then it should be stated; however, the mink eating only mussels represents an unreasonable scenario. On the other hand, this approach may be appropriate in the case of the bald eagle eating fish as a sole food source although a variety of species should be considered. Unless receptor organisms can be shown to preferentially rely upon a single food source, the single food item approach should not be used in screening or quantitative risk characterization.

3. Page 3, Section 4.1: This section indicates that in the risk characterization section of the risk assessment "a more reasonable maximum exposure ... will be estimated". Two points here: (1) where is the "risk characterization section of the risk assessment" (it does not appear to be presented in this document); and (2) not only should food sources be re-evaluated for "more reasonable maximum exposure", but other assumptions, equations, endpoints and benchmarks should be re-evaluated for the "risk characterization".. For example, NOAELs are appropriate for a screening- level assessment, but are inappropriate for "risk characterization" since NOAELs are measurements of no adverse effect, rather than quantifiable effects.

4. Page 3, Second paragraph (Section 4.1): It is noted that "until soil sampling is


conducted," it will be assumed that soil COEC concentrations are equal to those of sediment "in the various outfalls." The chemical state of many of the COECs will vary substantially between soil and sediment (hence the toxicity). Also, there is a significant "matrix effect" relative to ingestion of soil versus sediment. Depending upon the contaminant, this may be an ultra-conservative assumption. We disagree with using sediment concentrations for soils since soils data will be collected as part of sampling in the other site units. This approach will lead to misleading interpretation of results, particularly if ultra-conservative soil contaminant concentrations are used in conjunction with conservative soil ingestion rates.

It is also noted in the second paragraph of Section 4.1 that it will be assumed that "the ROCs use the outfalls as their only water source." This is not applicable for the river since the outfalls are being evaluated as part of the site assessment. In addition, from a technical standpoint, the water source should be based on consideration of receptor-specific home ranges.

5. Page 3: The remainder of Subsection 4.1 is a synopsis of exposure assumptions for each ROC (see also Table 2) for which there is a lack of documentation and associated literature citations to substantiate the statements. A few examples:

• bald eagles eat "particularly shad" This is likely to be a valid assumption, locally, and there are probably several corroborating literature references.

• what is the basis for stating that both the kestrel and red-winged blackbird are resident species commonly observed in MRP15? This may be the case, but it should be referenced with respect to credible literature/reports.

• it is stated that in the process of ingesting seeds and terrestrial invertebrates the red-winged blackbird is assumed to ingest sediment (this should be soil) at a specific rate. This is probably unlikely; we suspect that detailed behavioral literature will show that, other than in plowed fields, red-winged blackbirds seldom visit the ground per se and would be unlikely to ingest much soil or sediment near water bodies.

6. Page 3 (Raptor): "To be conservative, because of the limited information and because of the uncertainties inherent to specific BCFs (see Section 6.0), the highest BCF for each COEC will be applied to the bald eagle’s food source." We assume that the authors refer to the highest BCF reported in fish and not the "highest available BCF for each COEC." Please clarify. We also point out that in the case of PCBs, there is site-specific information already available on the concentrations in various fish species of MRP15 (YMA undated; WCC 1992; WCC 1994) as well as


mussels (USCOE 1984). These data should be used instead of estimating tissue concentrations using BCFs.

7. Page 3 (Raptor): "The American kestrel hunts both small mammals and small passerine birds. For the purpose of this assessment, the American kestrel will be modeled as consuming the white-footed mouse and, separately, passerine birds as represented by the red-winged blackbird." Since the guidelines assume both food sources are present at the site, it is illogical to then assume that the kestrels will not take advantage of whichever food source is more plentiful or simply easier to catch at various times of the year. As stated in comment 2, unless receptors can be shown to preferentially rely upon a single food source, the single food item approach should not be used in screening or quantitative risk characterization.

8. Page 4 (Shorebird): An inappropriate application of terminology is noted — namely, the application of the term shorebird to the great blue heron. There is a clear distinction between shorebirds and wading birds. Application of the term shorebirds is generally limited to the Suborder Charadrii (of Charadriiformes), which includes birds such as plovers, sandpipers, killdeer, etc. The herons, egrets, storks, etc., are members of the Order Ciconiiformes ... and not phylogenetically very close to the "shorebirds." The generalization that "Due to their habitat and eating habits, shorebirds also consume relatively high amounts of sediments" [no literature citation provided] is valid. Many true shorebirds feed by literally mining littoral and supralittoral sediments. However, it is incorrect to apply this to ciconiiform birds, particularly the great blue heron. There is a substantial amount of behavioral literature to show where and how various wading birds obtain food (e.g., Kushlan 1978). Flamingos, storks, and ibises (all ciconiiforms) do feed similarly to some of the charadriiformes — but virtually none of the ardeids (herons, egrets) do.

We recognize that "No ingestion rates specific to the great blue heron were available" and that a surrogate sediment ingestion rate was selected (i.e., from the sandpiper [a shorebird]). However, from an ecological standpoint, it is inappropriate to use the sandpiper as a surrogate for the great blue heron. Depending upon the basis for selection of ROCS, this is the type of information that can be used in a re-evaluation of ROCs as the ecological risk assessment process develops (see comment Selection of Receptors of Concern on page 5).

9. Page 5 (Reptile): "Due to a lack of toxicological data specific to reptiles, the aquatic and terrestrial reptile functional groups will not be evaluated quantitatively. Due to the uncertainties and assumptions already inherent in the baseline risk assessment (see Section 6.0), applying toxicological data from a different functional group would not be scientifically defensible." With the exception that the current screening-level approach is more qualitative than quantitative, we agree with this statement and the fact that extrapolating toxicological data from one functional group to another is not appropriate. This is corroborated by Opresko et al. (1994) and Suter et al. (1992)


which indicate that it is not considered appropriate to apply benchmarks across different groups. The Procedures document goes on, however, to develop ecotoxicological benchmarks for birds based on mammal data and application of an uncertainty factor of 10. It was not appropriate for reptiles, and it is not appropriate for birds. This has no basis in science because unrelated species have different toxicological responses to chemicals. A high degree of uncertainty would be associated with an ecotoxicological benchmark for an animal in a functional group that is different than the test species. Rather than developing a number that has very little meaning, the document should indicate that insufficient toxicity information is available to evaluate the bird functional groups for these particular chemicals.

10. Page 5: Subsection 4.2 discusses how direct exposures to strictly aquatic receptors will be evaluated. Essentially, this will be a desk-top exercise of comparing water and sediment concentrations of COECs to benchmarks such as AWQCs (which are referred to as "endpoints"). The equilibrium partitioning approach is intended to calculate "interstitial pore water" concentrations and then compare these to AWQCs on behalf of "the benthic feeders", mussels and aquatic invertebrates.

With respect to the organisms to be evaluated, only infaunal invertebrates (not all, and perhaps largely limited to oligochaetes and chironomids) are theoretically exposed to the pore water. Inclusion of benthic feeders, presumably referring to the river carpsucker, and aquatic invertebrates in general, is not appropriate. In addition, mussels respire by means of withdrawing water (from the water column) via their incurrent siphons, and probably do not normally draw from interstitial pore water. Therefore, comparing pore-water concentrations to AWQCs is probably inappropriate in the case of mussels as well.

With respect to the method for calculating pore water concentrations, the equilibrium partitioning approach is currently functional only for nonpolar organics (which, based on the COECs, would include the PCBs and PAHs; some volatile organics are nonpolar as well) and is not appropriate for polar organics or inorganics. The application of equilibrium partitioning theory to divalent cationic metals was scheduled for review by EPA’s Science Advisory Board in January of this year; the approach is currently not appropriate for use in applied scenarios.

An example of another potential method is the critical residue approach, which can be used with polar and nonpolar narcotics (e.g., some PAHs) (USAE 1992; McCarty 1991; McCarty 1986; Landrum 1992).

11. Page 5, last paragraph: Thomann and Connolly (1984) is referenced as the only study located which evaluated the effects of the ingestion of contaminated food on fish. Thomann and Connoly was a modeling manuscript and did not specifically address toxicological effects due to ingestion of contaminants. There are a number of relevant references in EPA’s Food and Gill Exchange of Toxic Substances User’s


Manual (Suarez and Barber 1992) that should be consulted on this matter.

12. Page 6: Per Subsection 5.1, exposure to all ROCs other than the strictly aquatic forms will be based on a calculated applied daily dose (ADD) - a combination of ingested food, water, and sediment/soil. Essentially the equation from Wildlife Exposure Factors Handbook (USEPA 1993) is adopted. This includes the expression F, the "diet fraction" (which is the percent of ingested media derived from the source of concern; that is, the fraction of the ingested media that is "contaminated"). Will the F term be developed as an area use factor, i.e., the area of contamination divided by the size of the home range (up to a value of 1)?

In addition, the equation has no accommodation for multiple food sources, nor are multiple food components with associated diet composition presented or even incorporated into the current approach. For most organisms, this is an unrealistic approach (see comment 2).

13. Pages 6 and 7: Beginning at the 3rd paragraph under 5.1 (bottom of page 6 and upper 3rd of page 7) is a discussion of "species-specific and contaminant-specific BCFs." It should be clarified in the beginning of this discussion that (we presume) these values are going to be used to estimate the concentrations of COECs in the various dietary components (CF^as mentioned briefly in Subsection 4.1). It would also be appropriate to provide the rationale and a general clarification of the approach; this section is, from the standpoint of an ingestion pathway, about the broader concept of bioaccumulation (uptake from food and water) rather than bioconcentration (uptake from water only).

In the last paragraph on page 6, the document indicates that "Bioaccumulation in terrestrial food chains was assumed to potentially occur for all organic COECs having octanol-water partitioning coefficients (log KJ of greater than 3.5." Please provide a reference to demonstrate the validity of this value for a terrestrial-based system. Will a BCF value of 1.0 be used for all organic COECs having a log K^s less than 3.5?

In the first full paragraph on page 7, an equation is presented for calculating BCFs in terrestrial vegetation. Note there are also equations for calculating BCFs for other organisms based on K^, and/or solubilities and lipid content that should be considered, e.g., Landrum (1992); USEPA (1991); USEPA (1994); Veith et al. (1980); and Chiou (1981).

In the first full paragraph on page 7, it is indicated that "the appropriate BCFs will be applied to the sediment concentration of each COEC". It is inappropriate to apply a BCF, a water-based factor, to a sediment concentration. There are biota-sediment accumulation factors (BSAFs) for some constituents which would be appropriate (e.g., Parkenton et al. 1992; USEPA 1994c).

Comments. EA \Mai95 13

In the discussion of BCFs, there is no indication as to which BCF value from Table 3 will be used when a range is presented, though on page 3 it is indicated that the maximum BCF is indicated for the bald eagle (see comment 6). When a range of values is available, we suggest the average reported BCF value be used in these calculations.

On the beginning of page 7, there is a statement that "if information regarding BCFs for a particular functional group cannot be located, BCFs from another functional group will be applied". It is inappropriate to apply BCFs across functional groups, just as it is inappropriate to extrapolate ecotoxicological factors across functional groups (see comment 9).

14. Page 7: An equation is presented with an “ingestion fraction” (IF) term. Though the derivation of the IF term can be deciphered from the values presented in Table 2, this should be explained in the text. It should also be explained how the result of this equation on page 7 relates to the calculation of the applied daily dose (ADD).

15. Page 7: Subsection 5.2 indicates that predator fish receive most of their exposure via ingestion of contaminated food, citing Thomann and Connolly (1984), which deals with lake trout and PCBs. As indicated previously, Thomann and Connolly was a modeling manuscript, and does not support the generalization that all predator fish receive most of their exposure through ingestion. It may or may not be valid for specific ROCs and COECs, but it has not been substantiated. Note also that because a chemical is bioaccumulated, it does not necessarily mean it poses a corresponding degree of hazard to the organism doing the accumulating.

16. Page 7, paragraph 5: "To be conservative, the highest concentration of the COECs in the surface water and sediment will be utilized in order to develop a ‘worst-case scenario’ of potential risks." Assuming continued exposure of an ecological receptor to maximum constituent concentrations is inappropriate even for a screening level assessment. The goal of the assessment is not to provide a worst-case evaluation of potential risks to individual receptors of concern, but rather to characterize potential risks to the ecosystem (i.e., population-level effects) in MRP15. The 95 percent upper confidence limit on the mean concentration (95 percent UCL) would be a more appropriate concentration to use in the comparison to ecotoxicological benchmarks. Populations are exposed to a range of concentrations which are best represented by the observed range of concentrations provided by the sampling results. This was discussed in regards to the human health assessment for the site and applies to the ecological assessment as well.

17. Page 8: Subsection 5.3 bears the title "Endpoints," which is extremely confusing and basically inconsistent with eco-risk guidance. In terms of more conventional terminology, this subsection is about ecotoxicological benchmarks, reference toxicity values (RTVs), toxicity reference values (TRVs), or simply effects levels. That is, the


discussion is about the levels (expressed as either doses [rates] or concentrations) that are known or expected to elicit adverse effects. The Framework uses the term endpoint in the context of assessment endpoints and measurement endpoints. The former are essentially expressions of overall goals or objectives (e.g., protection of a resource, a statement of a desired outcome), whereas the latter are absolute or relative indices for evaluation of fulfillment of the goals (e.g., population density). In using this terminology, we are concerned that if an exposure exceeds an ecotoxicological benchmark for the COEC/ROC combination, then it will be incorrectly concluded that a problem has been identified. In this sense the chosen effects level per se becomes something like what is termed a measurement endpoint in the Framework. The use of these values is appropriate for a screening level assessment, but the terminology should be modified to be consistent with conventional terminology.

18. Page 8: The 2nd paragraph under Section 5.3 talks about the sources of theecotoxicological benchmarks. The text alludes to the desirability of using original literature sources, but it is acknowledged that most of the numbers come from "profiles" (synoptic review articles). Reliance on the review literature ("profiles") is in opposition to current EPA guidance (USEPA 1992) and is inappropriate for risk characterization. See comment #28 for examples of problems associated with relying on profile documents.

19. Page 8 (and Tables 4 and 5): There appears to be little consistency in the selection process used to determine appropriate ecotoxicological benchmarks for the ROCs. Page 8 of the document provides the selection criteria used to develop the ecotoxicological benchmarks. The document indicates that the ideal scientific study was one that was conducted for a specific ROC with the objective of finding the No Observed Adverse Effects Level (NOAEL). If the selected value was not a NOAEL, an uncertainty factor of 10 was applied and if the selected value was applied to a different functional group, an uncertainty factor of 10 was applied. A review of Tables 4 and 5 indicates that in many cases, the “estimated values” provided in Suter et al. (1992) were used in the determination of the benchmark. The text does not discuss the methods used by Suter et al. to develop these estimated values. The scheme presented by Suter et al. for calculating estimated values is not consistent with the safety factors presented in the text, so when available, the toxicity values should be used to develop benchmark values, not the criteria derived by Suter et al. If these values are going to be used at all in the document the methodology used to develop the estimated values should be presented in the text.

In addition, for constituents that have promulgated criteria, the criteria should take precedent over the values developed by Suter et al. (except in the case of the Higginseye mussel). In the screening assessment, a selection process as follows should be used to ensure consistency and use of defensible data to develop ecotoxicological benchmarks for the ROCs:


1) State of Iowa Criteria;2) USEPA criteria;3) Chronic NOAELs (appropriate functional group);4) Chronic LOAELs (appropriate functional group);5) Non-chronic NOAELs;6) Non-chronic LOAELs;7) Acute Criteria;8) LDj0s and LC50s

This process is also consistent with the selection of COECs. Assuming the above scheme is used to develop ecotoxicological benchmarks for screening, the footnotes in Table 4 will need to be adjusted accordingly as well as the benchmarks presented in Table 5. In the case of the Higginseye mussel, specific toxicity data for mussels would be appropriate, if available.

20. Page 9, Section 5.3.1: It is noted that comparison of exposures and "endpoints" should "not be interpreted as a quantification of risk to ROCs". This is an important statement in the context of the screening level approach presented in this document, but, as already indicated, there is a lack of discussion regarding how the results will be interpreted, and what the next step is in the "risk characterization". The overall approach appears nearly identical to the so-called quotient method, but the hazard quotients and indices are not actually calculated (the ratio of the ADD to the ecotoxicological benchmark). We suggest the quotient method be used for screening purposes. Constituents that exceed a hazard quotient of 1.0 would be carried into the risk characterization. Those with a hazard quotient less than 1.0 would not longer be considered contaminants of ecological concern. (See comments regarding Application on page 6).

21. Page 9, Section 5.3.2: “Further research into the toxicity of contaminated sediments to mussels will be conducted in order to find values that more accurately represent the potential risk. ” Please provide additional information as to what this research will entail (e.g., database review?) and where the information will be documented once (if) it is found.

22. Section 6.0 is a discussion of uncertainties and data gaps. This discussion is thorough, but most of the uncertainty factors mentioned have to do with doubts about whether the value and/or approach chosen is sufficiently conservative. In a few cases it is acknowledged that the approach may lead to an overestimation of exposure and/or risk, but generally the potential for underestimation (i.e., Type II errors or "false negative" conclusions) is emphasized. Other sources of uncertainty include:

• Use of NOAELs (or estimated NOAELs) versus a dose/response relationship(e.g., a lowest available LOAEL) significantly contributes to uncertainty. Indeveloping empirically-measured NOAELs, the process is biased toward


selection of test concentrations rather than contaminant-associated effects. Use of NOAELs (measured or predicted) is appropriate for a screening-level risk assessment, but will not enable a quantitative evaluation of ecological risk.

• human error

• The assumption of simplified diets is a significant source of uncertainty. This is presumably done on the assumption that a consumer eating only one type of food (those assumed a priori to be the most contaminated) is "conservative". It is unclear how this approach will be used to interpret the results of the screening assessment (see comment #2). This approach is unrealistic in that virtually all higher consumers have extremely complex diets, within the constraints of their respective basic morphological/ physiological/behavioral adaptations.

• No mention is made of the uncertainty associated with assuming that assimilation via ingestion is 100% efficient. Assimilation cannot be complete per thermodynamic laws, not to mention the stochasticity in this efficiency, as would be introduced, for example, by different matrices, bioavailability and digestive limitations.

23. Page 10: Under the fourth bullet on page 10, the remark is made that the reported ingestion rates, body weights, and home ranges are "means for that species." Actually, this is incorrect and contradictory to earlier statements. As stated elsewhere in the document, in most cases the home ranges selected are the smallest "means" (if not extremes) found (e.g., American kestrel - "mean" territory size for adult males in California in winter; mallard - apparently the lowest extreme value for laying females in prairie potholes).

24. Table 1: Many VOCs are listed as COECs in surface water and sediment based on positive detections (rather than comparison to any standards or criteria). VOCs should not be construed as a concern in sediments unless there is a continuing source. Additional values of concern for surface water (e.g., LOELs) are presented in the NOAA quick reference cards (see comment 5 on the document entitled Preliminary Identification of Contaminants of Concern, Ecological Risk Assessment, Mississippi River Pool 15). These values should be used in the evaluation of COECs in surface

water.

25. Exposure Assumptions (Table 2)

There is an apparent oversight in the sediment/soil ingestion rates (column 5) provided within Table 2. Food ingestion rates were obtained primarily from EPA’s Wildlife Exposure Factors Handbook (EPA 1993) which are expressed as a fraction


of bodyweight per day (g/g-day) on a wet-weight basis. Soil/sediment ingestion rates are then expressed as a fraction of food ingestion rates on a dry weight basis (EPA 1993; Table 4-4, page 4-20). Thus, to estimate exposure to a wildlife species, either the food ingestion rate must be converted to a dry-weight basis or the sediment ingestion rate must be converted as a percentage of food intake on a wet-weight basis. EPA (1993) recommends that food ingestion rates be converted to a dry- weight basis when utilizing these combined exposure pathways (EPA 1993; Section 4.1.3.4, page 4-21). Thus, the values reported in Table 2 for the sediment/soil ingestion rate are significantly overestimated because a dry-weight fraction for sediment/soil ingestion rate was applied to the wet-weight food-ingestion rate for each species.

The amount of sediment/soil ingested by each of the eight bird and mammal ROCs referenced in Table 2 is based on an internal EPA document by Ruth Miller, "Survey of soil ingestion by wildlife", (November 1992). We are particularly interested in how the "conservative estimate based on available species" was derived. The data that went into these estimates should be provided in the text or in the table to give the reader an idea of how “conservative” the values are. In addition, the food, water and sediment/soil ingestion fractions are not well defined in the table or in the text.

Note: a complete citation should be provided for Peterson and Nebeker, 1992, "Estimation of waterborne selenium concentrations that are toxicological thresholds for wildlife."

.EA\Mar95 18

25. Exposure Assumptions (Table 2) (continued)

American kestrel

the "home range" value of 13 ha is based on the smallest area cited in USEPA (1993), which is the mean territory size for adult males in California in winter. There are other values available for the midwest which would be more applicable to the Davenport site, the lowest of which is 131 ha and is based on breeding pairs in summer in woodlands and fields of southern Michigan.

Bald eagle

"home range" is stated as a linear distance (3 km), with no reference to a source. Please cite the basis for the number and also explain how it will be used in estimating a diet fraction.

Red-winged blackbird

the sediment/soil ingestion rate of 0.00058 kg/day converts to about 5 % of the food ingestion rate, which is extraordinarily high for a bird that seldom visits (much less, feeds on) the ground except in plowed fields.

the "home range" of 0.81 ha is said to have been "extrapolated from robin data". How was the home range estimate extrapolated from one bird species to another, and why is the robin considered an appropriate species upon which to base the extrapolation?

Great blue heron

the sediment/soil ingestion rate of 0.0722 kg/day converts to 18% of the food ingestion rate (or 12.6% of all matter ingested), which is overly conservative for a bird with a diet and feeding behavior like the heron.

the "home range" is a very conservative size but supported by one credible set of observations. However, if the animal actually has a territory this small, then only one or very few individuals will be feeding in the contaminated area. This will impact interpretation of population-level effects of the great blue heron in MRP15.

25. Exposure Assumptions (Table 2) (continued)


Mallard

there is no reference for the "home range" of 38 ha; this number apparently came from USEPA (1993), where it is the lowest extreme value presented (for laying females in prairie potholes, whose mean size is actually 111, and 4 times that when they are not laying). The assumption presumes then that all mallards present are females, which are continuously brooding offspring. This is a good example of a value that could be easily verified in the field by a qualified observer. Though it might arguably be used for screening purposes, this value should not be used in the subsequent "risk characterization".

26. Table 3: The table indicates that many constituents (specifically VOCs) have low potential to bioaccumulate in invertebrates, mammals, and birds. Will the BCF in the equation on page 7 then be 1 for these constituents?

The text (page 7) acknowledges that a number of BCFs have not yet been identified for certain ROCs, and there is a notable lack of information regarding BCFs and BAFs in Table 3. BCFs and BAFs for many of these constituents can be found, for example, in ASTER (Assessment Tools for the Evaluation of Risk), OHMTADS (Office of Hazardous Materials Technical Assistance Database System), the Hazardous Substances Database and the open literature. Sources of BSAF compilations include Parkenton et al. (1993) and USEPA (1994).

27. Table 4: The units for the NOAA ER-L and ER-M values for PAHs and PCBs in sediment should be in parts per billion (ppb) rather than parts per million (ppm). Also, the source of these values referenced is the 1991 NOAA paper. The ER-L and ER-M values were recalculated by Long et al. (1993) after adding additional, more recent data and refining the approach. The revised NOAA screening values should be used for selection of COECs. Some of the values are higher than those presented in Table 4; others are lower. Based on the unit change and the updating of the NOAA reference, the following changes to Table 4 should be made:

Comments .EA\Mar95 20

CONSTITUENT NOAA ER-M (ppm)

NOAA ER-L (ppm)

Acenaphthene 650 to 0.500 150 to 0.160Anthracene 906 to 1.1 85 to 0.085Benzo(a)anthracene 1,600 to 1.6 230 to 0.261Benzo(a)pyrene 2,500 to 1.6 400 to 0.430Chromium 145 to 370 80 to 81Chrysene 2,800 to 2.8 400 to 0.384Copper 70 to 34 390 to 270Dibenz(a, h)anthracene 260 to 0.260 60 to 0.063Fluoranthene 3,600 to 5.1 600 to 0.600Fluorene 640 to 0.540 35 to 0.019Iron not reported in NOAA

tableLead 110 to 218 35 to 67Mercury 1.3 to 0.710 no change2-Methylnaphthalene 670 to 0.670 65 to 0.070Naphthalene 2,100 to 2.1 340 to 0.160PCBs 400 to 0.180 50 to 0.023Phenanthrene 13,800 to 1.5 225 to 0.160Pyrene 2,200 to 2.6 350 to 0.665Silver 2.2 to 3.7 no changeZinc 260 to 410 120 to 150

28. Tables 4/5: The values chosen as ecotoxicological benchmarks for the ecological risk assessment for some compounds may be inappropriate because 1) an intermediate reference was used rather than original source document to develop the criteria; and 2) attempts were made to extrapolate between functional groups which often leads to irrelevant and meaningless values. Examples are provided below and other examples may exist, highlighting the necessity for using original sources.

Table 4, page 11 of 39: For chromium (page 11 of 39), a chicken embryo study was used as the basis for the ecotoxicological benchmarks for raptors, passerines, shorebirds, and waterfowl. If the study was conducted by directly dosing chicken embryos with chromium, this LD50 would not be appropriate for derivation of ecotoxicological benchmarks for receptors at MRP 15. A more appropriate choice for the benchmark for raptors, passerines, and shorebirds would be the NOAEL of 100 mg/kg/day reported in the 32-day chicken study. Also, the reduced survival


benchmark of 10 mg/kg/day reported for American black ducks is a more relevant benchmark for waterfowl than is the concentration reported in the chicken embryo studies.

Ecotoxicological values for benzo(a)pyrene (BaP) for the white-footed mouse and mink were obtained from Opresko et al. (1993). This document reported a toxicological value of 10 mg/kg for laboratory mice based on a study from Mackenzie and Angevine (1981). Mackenzie and Angevine noted potential reproductive effects (i.e., decreased fertility in the FI progeny) in laboratory mice (strain CD-I) at a LOAEL of 10 mg/kg/day. The dose was administered by gavage. In contrast to this study, BaP administered in food demonstrated no adverse effect in the fertility of Swiss mice (Rigdon and Neal 1965) suggesting that bioavailability of BaP may be limited when administered in food versus that obtained orally but artificially via gavage. Thus the applicability of this study (Mackenzie and Angevine 1981) in the development of a benchmark for this assessment may be considered questionable although the assessment endpoint (potential reproductive / developmental effects) is not questioned. Note that a number of values reported by Opresko et al. (1993), including BaP, have since been revised (see comment 38).

To summarize, ecotoxicological values, if obtained from the literature, must be obtained from the original source and not from an intermediate reference. Care must be taken to obtain revisions to source documents to assure accuracy. Extrapolations should only be made between species of a similar functional group using assessment endpoints that are relevant for the species of concern. Therefore, assessment endpoints for a particular functional group should be based on all available studies on a particular chemical but should consider only data that are relevant for a functional group. However, measurement endpoints should be based on a subset of studies that are most appropriate for evaluating potential impacts from the exposure pathways of concern.

Finally, measurement endpoints should be used within the context for which they were developed. For example, Opresko et al. (1993) states that the values proposed were to be used in a tiered approach - Tier I for a screening level assessment (i.e., to identify concentrations that are nonhazardous) and/or Tier II as a part of a weight- of-evidence approach considering multiple lines of evidence to identify potential ecological effects. Thus, the values proposed in Table 5 could not be used alone to identify potential ecological risk and must consider other lines of evidence.

29. Table 4, Page 32 of 39 (tetrachloroethylene) and page 35 of 39 (trichloroethylene): The acute and chronic criteria reported are actually LOELs. The effects columns should be adjusted accordingly.

30. Table 4: There are some missing literature citations given for the "toxicological values" (e.g., Marchini et al. 1992; Zoto and Robinson 1985).


31. Table 5: An ecotoxicological benchmark of 0.012 /xg/L mercury is used for fish and mussels. This is based on the chronic ambient water quality criterion. However, the chronic water quality criterion is based on human health rather than ecotoxicity (final residue value). An appropriate ecologically-based effect (growth, survival, reproduction, etc.) of mercury would be different (see comment 38: mercury).

32. Table 5: For copper, the proposed benchmarks for fish and invertebrates are below both the national ambient water quality standard and the Iowa water quality standards, both of which are judged by the respective agencies to be sufficiently protective of aquatic biota. The proposed values may be appropriate in a screening approach, but are not appropriate in the final risk characterization. An exception to this is that a value lower than the water quality standards may be appropriate for mussels, since Lampsilis higginsi represents a local species of special concern based on its endangered status. In this case, a value representative of the mussel (if available) should be used.

33. Table 5 has a column for “other” safety factors to be used in the development of toxic benchmarks. The use of these safety factors is not sufficiently documented in either the text or the table footnotes. The safety factors should be removed.

34. Table 5: It is inappropriate to apply an uncertainty factor of 10 when using rat data to develop toxic benchmarks for mink and opossum. Rats are omnivores; therefore, the functional group uncertainty factor of 10 should not be necessary.

35. Table 5, page 17 of 25 (2-methylnapthalene): A benchmark of 5 mg/kg/day is listed for all ROCs. This was then converted to mg/kg/bodyweight using appropriate adjustment factors. Was the original study reported in mg/kg bodyweight/day? We were not able to consult the original reference (Sandmeyer 1981) because it was not listed in the references.

36. Table 4: EPA’s Ambient Water Quality Criteria are provided in Table 4; however, Iowa’s water quality standards, which should be applicable over EPA’s criteria, are not included. Examples of differences in chronic criteria between Iowa and EPA are for mercury (0.05 /xg/L vs 0.012 /xg/L) and cyanide (10 /xg/L vs 5.2 /xg/L). (Please refer to Table 1 of Preliminary Identification of Contaminants of Concern, Ecological Risk Assessment, Mississippi River Pool 15).

37. Tables 4/5: Ecological risk is a rapidly evolving field, and information is being continually updated (e.g., the NOAA information °in Comment 27). For example, a number of references from Oak Ridge National Laboratory have recently been published, e.g., Opresko et al. (1994); Will and Suter (1994a); Will and Suter (1994b); Sample and Suter (1994); Suter and Mabrey (1994); and Hull and Suter (1994). Some of these references contain information that supersedes previously

published information that has been used in this procedure.


Many of the Suter et al. advisory values, which are referenced throughout Table 4, and are recommended values for a number of the constituents in Table 5, are no longer applicable. Publication of an updated toxicological benchmark report by Suter and Mabrey (1994) indicates that it "supersedes a prior aquatic benchmarks report (Suter et al. 1992). It adds three new types of benchmarks and deletes one that proved to be unreasonably conservative [the advisory values]. It also updates the benchmark values where appropriate, adds some new benchmark values, replaces most secondary sources with primary sources, and provides more complete documentation of the sources and derivation of all values." Some of the revised values are provided below:

CONSTITUENT Suter et al. 1992 (/ig/L)

Suter and Mabrey 1994(lig/L)

Benzo(a)anthracene 0.0027 0.027Benzo(a)pyrene 0.0013 0.0142-butanone acute: 372,000

chronic: 20,800Carbon disulfide acute: 159

chronic: 8.89Dibenzofuran 2.0 20.4Manganese 11 80.3Methylene chloride 220 2,240Phenanthrene 8.4 3.23Toluene 10.4 1761,1,1 -T richloroethane 1.4 62.1Vinyl chloride 0.95 87.8Xylenes 1.46 86.2

In addition, some of the values in Table 4 are inaccurate as reported (e.g., bluegill LC50 for acetone). Comments on specific constituents are provided below:

Commcnts.EA\Mar95 24

Acetone: The number for bluegill in Table 4 (and 5) for is incorrectly reported. Verschueren (1983) reports an LC50 for bluegill of 8,300 mg/L. Note also that Verschueren (Handbook of Environmental Data on Organic Chemicals) is a compilation of data, and the original reference is not provided in the Procedures document. There is an incomplete citation for Cowgill and Milazzo (1991) referenced for Daphnia magna toxicity (there are other citations missing as well, e.g., Clarke and Clarke [1967] referenced for zinc).

Aluminum: A safety factor of 10 is applied for fish because the reported value is not a chronic value. However, the value is referenced as a 7-day exposure (complete citation not provided), which is considered a chronic exposure based on EPA shortterm chronic toxicity testing protocols. A safety factor of 10 is not appropriate. The ambient water quality criterion for aluminum is 748 /zg/L (EPA 440/5-86-008, 1988), which is an appropriate example of an ecological benchmark for aquatic receptors, perhaps with the exception of mussels due to the endangered status of Lampsilis higginsi.

Benzene: The lowest chronic value for fish in Suter and Mabrey (1994) is 8,250 /zg/L as opposed to 299 /zg/L reported in Table 4. An estimated chronic value is 45.5 /zg/L; the population EC20 reported is 229 /zg/L. The LC50 value reported in Table 4 and Table 5 of 0.025 /zg/L for juvenile fathead minnows is completely inconsistent with the literature for the toxicity of benzene and should not be used.

Benzo(k)fluoranthene: The value presented is a human health-based criterion and is inappropriate for evaluation ecological risk.

Benzo(g,h,i)peryIene: The 0.0311 /zg/L value reported in Table 4 is a human-health based concentration and should not be used to assess ecological risk.

Chromium: the benchmark listed is the NAWQC chronic criterion for chromium VI (11 /zg/L). It is unlikely that chromium is present in the hexavalent state; this may be an item that would be easily measured during field studies if the screening process does not eliminate chromium as a COEC. Note that the population EC20 proposed by Suter and Mabrey (1994) is 316 /zg/L.

Dibenzofuran: 2.0 /zg/L represents an unrealistic aquatic life benchmark for fish. USEPA ASTER (QSAR) (1994) calculates the LOAEL for fish at 104 /zg/L and for macroinvertebrates at 148 /zg/L.

Fluoranthene: The lowest chronic value reported by Suter and Mabrey (1994) for

daphnids is 15 /zg/L, which is lower than that reported in Table 4. The lowest chronic value reported for fish is 30 /zg/L, which is not included in Table 4.

Iron: Why are uncertainty factors applied to the aquatic receptors in Table 5? The


chronic water quality criterion for iron is 1,000 /zg/L.

Mercury: The proposed values for mercury are different than the referenced values for Suter & Mabrey (1994). For example, the population EC20 reported by Suter and Mabrey (1994) for inorganic mercury is 0.32 /zg/L and the calculated chronic value is 1.3 /zg/L. These values are 0.28 /zg/L and 0.003 /zg/L, respectively, for methyl mercury. The lowest reported EC20 values for fish and daphnids are < 0.03 and 0.87 /zg/L for methyl mercury; and 0.87 /zg/L inorganic mercury for both fish and daphnids. EPA’s ambient chronic water criterion of 0.012 /zg/L is human health- based (a final residue value based on methyl mercury) and is inappropriate to use for assessing ecological risk.

4-MethylphenoI: The Handbook of Environmental Data on Organic ChemicalsVerscheuren (1983) presents only a compilation of data. The original source references should be provided.

Pyrene: the QSAR (quantitative structure activity relationship)-derived MATC(maximum acceptable toxicant concentration, now referred to as the chronic value) is 61/zg/L vs the 100 /zg/L listed in Table 5.

38. Table 4/5: The toxicological benchmarks in the 1994 revision of Opresko et al. for wildlife diets are generally higher than those listed in Tables 4 and 5. Some specific examples are:

CONSTITUENT Opresko et al. 1992 Opresko et al. 1994

MinkWhitefootedMouse

MinkWhitefootedMouse

Benzene 1.5 6.4 09 20

Benzo(a)pyrene 0.0028 0.0011 2.3 7.17

Cyanide 6.7 27.8 32.67 102.06

1,2-Dichloroethene 24.1 100.3 103.72 323.99

Manganese 0.50 2.07 454.26 1,418.96Methylene Chloride 3.62 15.04 30.20 94.33

PCBs (Aroclor 1254) 0.07 1Tetrachloroethylene 8.3 34.3 3.21 10.04

Toluene 13.8 57.3 59.62 186.22

Trichloroethylene 46 193 2414.45 7541.99

Vinyl Chloride 0.08 0.33 0.88 2.74

Xylenes 310 1,286 4.73 14.77

Zinc 6.0 24.9 825.93 2,579.92

Comments .EA\Mar95 26

4.

Methylene chloride: The toxicological benchmark NOAEL from Opresko et al. (1994) is reported 25.19 mg/kg for red fox, as opposed to 2.29 mg/kg in Table 4.

4-Methylphenol: The Handbook of Environmental Data on Organic ChemicalsVerscheuren (1983) presents only a compilation of data. The original source references should be provided.

PCBs (Aroclor 1254): Homshaw et al. (1983) is referenced as proposing a chronic PCB criterion for mink. The basis for Homshaw’s value is unknown, and it is inconsistent with other reported values (based on the reported units). This value should not be included in Table 4.


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Landrum, P.F., H.Lee II and M.J. Lydy. 1992. Toxicokinetics in Aquatic Systems: Model Comparisons and Use in Hazard Assessment. Environmental Toxicology and Chemistry, Vol. 11. pp. 1709-1725.

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McCarty, L.S. 1991. Toxicant Body Residues: Implications for Aquatic Bioassays with Some Organic Chemicals. Aquatic Toxicology and Risk Assessment: Fourteenth Volume. ASTM STP 1124, M.A. Mayes and M.G. Barron. Eds. American Society for Testing and Materials, Philadelphia, pp. 183-192.

McCarty, L.S. 1986. The Relationship Between Aquatic Toxicity QSARS and Bioconcentration for Some Organic Chemicals. Environmental Toxicology and Chemistry, Vol. 5, pp. 1071-1080.

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National Oceanic and Atmospheric Administration (NOAA), 1991, “Potential for Biological Effects of Sediment-Sorbed Contaminants Tested in the National Status and Trends Program, Memorandum NOS/OMA 52, Long E.R. and Morgan, L.G.

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Parkenton, T.F., J.P. Connolly, R.V. Thomann and C.G. Uchrin. 1993. Do Aquatic Effects or Human Health End Points Govern the Development of Sediment-Quality Criteria for Nonionic Organic Chemicals.

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letter transmitting comments regarding the preliminary ... · 1. page 1: paragraph 2 and table 4....

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