### September 1, 2009



# Responses to Review Comments on Human and Ecological Risk Assessment for Coal Combustion Wastes

Final Draft

Submitted to:

Becky Cuthbertson Office of Resource Conservation and Recovery Office of Solid Waste and Emergency Response 

### U.S. Environmental Protection Agency

Submitted by:

RTI International

EPA Work Assignment 0-02 EPA Contract No. EP-W-09-004DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

iii 

### Table of Contents

1.0 Introduction and Background ................................................................................................... .............. 1

2.0 Objectives and Organization ................................................................................................... ............... 1

3.0 Comments on Groundwater Modeling .............................................................................................. ..... 1 3.1 10,000-Year Risk Duration .................................................................................................... ..... 2 3.2 Groundwater Transport Mo deling Approach .............................................................................. 4 3.3 Site-Specific Analysis and Comparisons with Measured Data ................................................. 10 3.4 Well-Related Issues ........................................................................................................... ........ 14

4.0 Source Modeling ............................................................................................................... ................... 19 4.1 Use of TCLP Data and the Availa bility of the Vanderbilt Study .............................................. 19 4.2 2006 DOE Facility Data a nd Liner Scenarios ........................................................................... 23 4.3 Alternative Waste Disposa l/Codisposal Scenarios .................................................................... 25 4.4 Liner Leakage ................................................................................................................. ........... 26 4.5 Treatment of Nondetect Values ................................................................................................. 29 4.6 Effect of Waste Compaction .................................................................................................... .. 31 4.7 HELP Infiltration Rates/Cover Assumptions ............................................................................ 32 4.8 Scenarios Where the Water Table Is Above the Disposal Unit ................................................. 32 4.9 Impact of Increased Mercury Capture with New Future Air Emission Control Regulations ................................................................................................................... ............. 33 4.10 Use of Porewater Data for Surface Impoundment Source Concentrations ............................... 34 4.11 Modeling Surface Impoundments Only Before Closure ........................................................... 35 4.12 Modeling Landfills Only After Closure .................................................................................... 36

5.0 Comments Related to Ex posure Scenarios ........................................................................................ ... 37 5.1 Direct Surface Impoundment Exposure..................................................................................... 37 5.2 Indirect Exposure Through Game Species (e.g., Deer, Waterfowl) .......................................... 39 5.3 Exposure to Surface Impoundment Effluent ............................................................................. 40 5.4 Exposure During Landfill Operation (Before Closure) ............................................................. 41 5.5 Elimination of Pathways During Screening Analysis ............................................................... 42

6.0 Comments Related to Human Exposure and Toxicity ......................................................................... 43 6.1 Combined Chemical Effects, Additive Risk, and Multiple Pathway Exposures ....................... 43 6.2 Toxicity Values ............................................................................................................... .......... 46 6.3 Target Cancer Risk ............................................................................................................ ........ 48 6.4 Defining the Exposed Population .............................................................................................. 48 6.5 Exposure Factor Assumptions ................................................................................................... 49 6.6 Child/Infant Exposure Assumptions .......................................................................................... 50 6.7 Correlations in Input Variables ............................................................................................... ... 52

7.0 Comments Related to Ecological Exposure and Toxicity .................................................................... 53 7.1 Cumulative Ecological Risk .................................................................................................... .. 53 7.2 Sensitive Habitats/S ensitive Species ......................................................................................... 54 7.3 Indirect Effects .............................................................................................................. ............ 54 7.4 Ecological Benchmarks, Ecological Toxicity ............................................................................ 55

8.0 Comments About the Monte Carlo Analysis Approach ....................................................................... 57 8.1 Use of Conservative Pa rameter Estimates ................................................................................. 57 8.2 Variability versus Uncertainty ................................................................................................ ... 58 8.3 Calculation of Percentile Probabilities as Opposed to an Overall Probability .......................... 59 8.4 The Numerical Stability of Monte Carlo Results ...................................................................... 59DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

iv 8.5 Impacts on Groundwater as a Resource..................................................................................... 60

9.0 Comments Related to the Documentat ion and Miscellaneous Comments ........................................... 62 9.1 Insufficient Documentation to Reproduce the Analysis Independently .................................... 62 9.2 Graphical Input/Output Distributions, Truncation Values ........................................................ 63 9.3 Documentation of Sensitivity Analysis ..................................................................................... 64 9.4 Discussion of General Risk-Susceptibility Patte rns such as Climate-Specific Risk ................. 64 9.5 Discussion of Worst-Case Scenarios ......................................................................................... 65 9.6 Discussion of Sulfate, Phosphate, and Chloride ........................................................................ 66 9.7 Background Concentrations ..................................................................................................... . 66

10.0 References .................................................................................................................... ............. 67DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

1 

### 1.0 INTRODUCTION AND BACKGROUND On August 29, 2007, the U.S. Environmental Protection Agency (EPA) published a Notice of Data Availability (NODA) to seek additional informa tion from the public on coal combustion waste (CCW) material to help inform its regulatory determination (72 FR 49714). To ensure protection of human health and the environment, EPA completed a risk assessme nt to evaluate the potential exposure to CCW in landfills and surface impoundments.

As stated in the NODA, the risk assessment was indepe ndently peer reviewed. A peer review is a process for enhancing a scientific or technical work product so that the decision or position taken by the Agency based on that product has a sound, credible basis.

The Agency contracted with IEc, Incorporated, to conduct a letter peer review of the draft Human and Ecological Risk Assessment of Coal Combustion Wastes.

This document provides responses to peer review comments received on the CCW risk assessment. There were five peer reviewers with the following focus areas:



### ▪ Dr. Nicholas Basta, Ohio State University: overall risk assessment approach, including the use of methods and parameters u sed in geochemical modeling



### ▪ Dr. Charles Harvey, Massachuse tts Institute of Technology: use of probabilistic approaches used in the analysis in the context of fate a nd transport of contaminants in the environment



### ▪ Dr. William Hopkins, Virginia Polytec hnic Institute and State University: ecological impacts as they were addressed in the analysis



### ▪ Dr. Kerry Rowe, Queens University: relationship between the use of landfill liners and the analytical approach



### ▪ Dr. Donna Vorhees, The Science Collaborative: adequacy of the human health risk assessment approach in the context of requirements for proper conduct of probabilistic risk assessments.

EPA has established a docket for this action (Doc ket ID No. EPA-HQ-RCRA-2006-0796). All documents in the docket are listed on www.regulations.gov . The public can view the peer reviewers’ comments in their entirety on the docket Web site. The docket is located at the EPA Docket Center, Public Reading Room, EPA West Building, 1301 Constitu tion Avenue, NW, Washington, D.C.



### 2.0 OBJECTIVES AND ORGANIZATION This comment review and response report focuses on the technical themes presented by the peer reviewers of the CCW risk assessment. For each comment category, the comments are provided by reviewer, followed by the response.



### 3.0 COMMENTS ON GROUNDWATER MODELING The peer review comments relating to modeling of the groundwater pathway are grouped into four subcategories:



### ▪ The 10,000-year maximum groundwater simulation time



### ▪ Groundwater transport modeling assumptions



### ▪ The need for more site-specific analysis and comparisons with measured data



### ▪ The number of wells contaminated, realistic risk of exposure, well placement within the plume, and other well-related issues.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

2 

### 3.1 10,000-Year Risk Duration

Comment Summary Both Dr. Basta and Dr. Rowe questioned the use of a risk assessment model that assumes current conditions will be maintained for 10,000 years and wondered how realistic the results may be. Specifically, Dr. Basta requested that the limitations of this assumption be better explained and questioned the assumption that contaminant concentrations in the leachate remain constant throughout the tim eframe. Dr. Rowe questioned the assumption that well use and climate conditions will remain constant for 10,000 years.

Dr. Nicholas Basta, Ohio State University
One limitation that is not discussed is the validity of modeling ri sk over a 10,000-year period. Shifting land use, geological and cl imate phenomena, and human and ecological populations will dramatically change exposure scenarios for CCW from surface impoundments and landfills. The time period for modeling risk should be discussed as a limitation and justified.

The assumption made that the contaminant c oncentration in the leachate data remains constant throughout the 10,000-year modeli ng period is not realistic. Leaching of contaminants will likely occur in much shorter time periods and decrease after reaching a maximum. The peak contaminant concentrati on will not remain constant throughout the modeling period. The assumption that 100% of the contaminant mass will be leached out of the CCW is unlikely. Only a fraction of the contaminant will leach from a solid waste such as CCW. The remainder of the c ontaminant will remain insoluble and “unleachable.”

Dr. Kerry Rowe, Queens University It appears to be assumed that the current hydrogeologic and climatic conditions will prevail over the next 10,000 years. While it is difficult to do otherwise, the implications of this assumption have not been adequately discussed. For example, it is assumed that well usage will correspond to that which exists today down gradient of Subtitle D MSW landfills (RA pg. 2-11) however global warming may result in the need for much greater reliance than at present on groundwater resources; if that were to be the case then the risks may be (considerably) higher than implied by the present study in even 30-100 years from now – let alone in 10,000 years time! The issue of time is also discussed in response to Question # 8.

The time period examined (10,000 years) is VERY long for the surficial hydrogeologic environments being considered – it is only about 10,000 years ago that glaciers retreated from much of Canada and northern USA. Thus the landscape that we see today is very different from that 10,000 years ago and we can expect it to be very different 10,000 years from now. Also the climate is likely to ch ange substantially over this period. Thus considerable skepticism needs to be applied to predicted impacts that are thousands of years from now – however it is not clear which results fall into this category (except for composite–lined sites where it is indicated that many impacts had not occurred in 10,000 years (RA, pg. 4-7)).

EPA Response EPA acknowledges that the 10,000-year modeling time hor izon should be further clarified in the revised risk assessment. The choice of a maximum 10,000-year time horizon does not mean that all simulationsDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

3 continue for the full 10,000 years. In the 2007 risk assessment, Section 2.4.2, page 2-15, it states “[t]he groundwater model is run until contaminant concentrati ons at the receptor point return to zero after the concentration peak or for the maximum simulati on time of 10,000 years, whichever comes first.” Although groundwater concentration predictions ar e simulated after the p eak or maximum average concentration is observed, the post-peak or post-ma ximum average predictions are not used in estimates of risk.

In many cases the leachate plume reaches the receptor point much sooner than 10,000 years for surface impoundments. Table 4-7 and discussions in Section 4.1.1. 5 of the revised risk assessment show that for unlined and clay lined surface impoundments, 70 percent of peak concentrations for arsenic III arrive at the receptor point within 91 and 150 years, resp ectively. Timeframes for landfills are significantly longer—unlined and clay lined units, 70 percent of peak concentrations for arsenic III arrive at 6,400 to >10,000 years. Composite liners also show much higher peak arrival times, close to 10,000 years for most surface impoundment runs and over 10,000 years for landfills. Under such timeframes EPA acknowledges that surface conditions may change si gnificantly, compounding the uncertainty associated with the predicted exposures and risks.

EPA acknowledges that the leaching profile described by Dr. Basta may be more realistic. However, the Agency does not have the data to use time-varying le achate concentrations. Therefore, the nonlinear fate and transport solution used for metallic constituents in the unsaturated zone module of EPA’s Composite Model for Leachate Migration with Transformati on Products (EPACMTP) is based on the assumption that the leachate concentration released from the waste management unit is constant over time (see Section 3.3.5.3, U.S. EPA, 2003b).

EPA acknowledges that there may be insoluble or ot herwise unleachable contaminant mass that remains in the waste management unit. However, there were no data available at the time of the 2003 risk assessment (including the Vanderbilt study) to suppor t a different approach. Thus, for this draft risk assessment, the leaching duration was assumed to be fin ite (see Section 3.4.2 of the draft risk assessment, Equation 3-1, and Section 3.4.2.2 for discussion) , and the waste concentration model input was conservatively assumed to be 100% of the constituent in the waste.

EPA acknowledges that future groundwater use pattern s may shift the number and location of receptors. However, there were no data available at the time the risk assessment was being conducted to support a different approach. The Agency used the best available data at the time , as discussed in Appendix C of the 2007 draft risk assessment.

A discussion has been added to the scen ario uncertainty section of the risk assessment (Section 4.4.1) that acknowledges potential limitations and uncertainties in risk estimates due to changing conditions over time, along with clarification of the time frame for the risk estimates in the risk assessment:

“The risk assessment assumed that contamin ant concentrations in the leachate remain constant throughout the 10,000-year modeli ng timeframe, although leaching may or may not persist for 10,000 years, depending on model inputs. The waste concentration model input was assumed to be a portion of the to tal waste concentration available to be leached, and it was assumed that 100% of th e constituent in the waste could leach from the landfill. The nonlinear fate and transport so lution used for metallic constituents in the unsaturated zone module of EPACMTP is based on the assumption that the leachate concentration released from the waste manage ment unit is constant over time (see Section 3.3.5.3 of U.S. EPA, 2003b). Although a leach ing profile that changes over time might be more realistic, the simplified leaching profile used by the model does not directly or indirectly translate into a poorer estimate of risk associated with groundwater exposures.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

4 The adoption of a simplified leaching profile to support a non-linear sorption approach in the unsaturated zone offered a greater bene fit and defensibility to the overall approach than assuming linear partitioning and a de pleting leachate profile would have.”


### 3.2 Groundwater Transport Modeling Approach

Comment Summary Comments related to the specifics of the groundwater transport modeling were received from Dr. Basta and Dr. Harvey. Issues covered in their comments included the following:



### ▪ The way that aquifer Kd values were dete rmined and used, including the fact that the risk assessment did not model precipitation-dissolution processes that probably control several of the CCW chemical constituents



### ▪ The approach used to determine the valu e of pH in the aquifer for selecting Kd



### ▪ The subsequent calculation of the retardation factor.

The comments also addressed the general need for more transparency in the data and methods used in the analysis and the need for validation and/or comparison of model inputs and results to site-specific field data.

Dr. Nicholas Basta, Ohio State University Another major source of uncertainty is th e ability of the MINTEQA2 model-generated Kd values used in EPACMTP to determine rea listic contaminant transport at field CCW facilities. Kd values for arsenic and other select contaminants should be determined for a set of soils with a range of soil chemical pr operties critical to chemical processes values. This could be conducted using column st udies under controlled laboratory conditions. Modeled Kd values used in the Draft must be validated by comparison to Kd values derived from real soil and/or geomedia to assure that the EPACMTP model captures the major chemical process critical to transport. Realistic Kd values will greatly affect the transport and risk calculation of As in groundwater. Kd values from soil or geomedia from CCW sites with a range of chemical properties will significantly reduce uncertainty associated with risk from As in groundwater (a risk driver) and significantly improve the risk assessment.

In general, all of the chemical processes ar e lumped into one value, Kd, calculated by MINTEQA2. This is a common modeling appro ach, although this grossly oversimplifies and is likely not an accurate approach for several CCW chemical constituents that are controlled by precipitation-dissolution pro cesses in soil that cannot be accurately modeled using Kd values.

Chemical processes are main ly described through determination of a partitioning coefficient, Kd, from nonlinear isotherms for each contaminant. The generation and use of an accurate Kd value is absolutely critical because it controls the amount of contaminant that is dissolved in the leach ate and capable of being transported. The contaminant is not transported if it is not di ssolved. Kd is very sensitive to adsorbents selected for inclusion in MINTEQA2. The type of sorbent varies with contaminant. In general, soil organic matter and clay minerals (especially Fe, Al, Mn oxides) sorb metal cations and oxyanions (i.e., arsenate and arsen ite). Soils vary greatly in the type and amount of these sorbents. The MINTEQA2 mode l is very limited in its ability to model sorption. MINTEQA2’s database is limited to three sorbents: Fe sorbent as the mineralDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

5 goethite, Fe sorbent as hydrous ferric oxide, a nd organic matter. Fe sorbent as goethite and organic matter were selected to calcula te nonlinear isotherms and subsequent Kd values in the Draft. The choice of Fe sorb ent is important because goethite is a much poorer adsorbent than hydrous ferric oxide a nd will result in larger leachate contaminant concentration. No explanation for the choice of goethite over hydrous ferric oxide is provided in the Draft. In addition to the type of adsorbent, the amount of adsorbent (i.e., adsorption sites) will greatly affect the calcul ated Kd value. The adsorption site density for the two sorbents is not specified in the draft. This is critical because we do not know if the sorption site density represents the normal range found in soils.

It is not possible to determine if the appr oach used by MINTE QA2 provides Kd values for the EPACMTP model that is representative of chemical processes in soils or geomedia at any of the CCW locations from information presented in the Draft. Soil has many more sorbents than the two sorbents used in the MINTEQA2 model in the Draft. Except for organic matter content, soil sorbents were not measured in soils collected from CCW facilities sites. The only soil data collected and reported was pH, organic carbon content, and soil texture. However, very im portant sorption of hydrous Fe, Al, and Mn oxides soil minerals were not reported.

Arsenic solubility and transport in soil is gr eatly affected by soil sorbents and Kd inputs into transport models. It is well known that the major chemical process that affects transport of As is adsorption of As(V) or As(III) to hydrous Fe, Al, and Mn oxide minerals. The amount of As transported w ill depend on the type and amount of these minerals in soil. These mineral sorbent phases were not measured in the soils from CCW facility sites and most are not in the MINTEQA2 model. Because one of the most critical chemical processes (sorption to oxides) affecting As transport was not used to calculate Kd values for input into EPACMTP, As solubility and subsequent transport is overestimated. Soil content of hydrous Fe, Al, and Mn oxide minerals ranges from < 0.5% for sandy soils to > 50% for weathered clay soils of the southeastern U.S., Hawaii, and Puerto Rico. The overestimation of As tran sport in soils with m oderate to high levels of hydrous Fe, Al, and Mn oxides could be orders of magnitude.

However, some insight into impact of excl usion of hydrous Fe, Al, and Mn oxides on As in leachate is provided from the Draft: “Sensitiv ity analysis results indicate that for most of the scenarios evaluated (over 75 percent), the risk assessment model was most sensitive to parameters related to groundwate r flow and transport: WMU infiltration rate, leachate concentration, and aquifer hydrau lic conductivity and gradient. For strongly sorbing contaminants (such as lead and cadmium ), variables related to sorption and travel time (adsorption coefficient, depth to groundwater, receptor well distance) are most important.” This statement implies that As is not a strongly sorbing contaminant in the EPACMTP model. Arsenic is strongly adsorb ed in soil and environmental media. Arsenic is more strongly adsorbed than those identified by EPACMTP as “strongly sorbing contaminants (such as lead and cadmium).” A likely reason for the poor sensitivity of the EPACMTP model for arsenic sorption is an inaccurate input Kd value generated from MINTEQA2.

One approach to answer the question posed, “Do the models capture the major physical and chemical processes and mass transport mechanisms that are important for the chemical contaminants and exposure scenari os studied?” is comparing Kd used for transport modeling using EPACMTP in the Dr aft with Kd values for select soils from CCW facilities. Kd values for As and other sel ect contaminants shoul d be determined forDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

6 a set of soils with a range of soil chemical properties critical to chemical processes values. This could be c onducted using column studies under controlled laboratory conditions. Modeled Kd values used in the Draf t must be validated by comparison to Kd values derived from real soil and/or geomedia to assure that the EPACMTP model captures the major chemical pr ocess critical to transport.

Arsenic speciation was considered in the Dr aft as a source of uncertainty. In the Draft (p. 4-30):

“Because the soil and groundwater models assume one form for each model run, the risk results presented for arsenic are based on arsen ic(III), which is a conservative, protective assumption (i.e., arsenic(III) has higher risks than arsenic(V)). Although arsenic is generally thought to occur in the +3 form in leachate, there is evidence from damage cases at CCW disposal sites that suggests that arsenic(III) is rapidly converted to arsenic(V) during subsurface transport, with th e result that drinking water standards are rarely exceeded in offsite groundwater in sp ite of high landfill leachate concentrations (see, for example, U.S. EPA, 2000; U.S. EPA, 2003e; Lang and Schlictmann, 2004; Zillmer and Fauble, 2004). To address this uncertainty (i.e., how much an arsenic III assumption might overpredict offsite well con centrations) the models were run assuming arsenic V as the arsenic species in soil and groundwater. Figure 4-4 compares the risk results for arsenic III and arsenic V. Arsenic V has lower risks than arsenic III by about a factor of two, As(III) vs. As(V)...”

The scientific literature has many examples showing As(V) has much lower solubility and mobility than As(III) in soil and environm ental media. Arsenic(V) which occurs as arsenate, is strongly adsorbed to clay / oxide fractions in soil and geomedia, whereas As(III) has a much lower adsorption capacity. Also, As(V) is more likely to form insoluble precipitates (i.e., scorodite) than As(III). Thus, As(V) is often magnitudes less soluble than As(III). It is striking that the differences in modeled risk reported in the Draft were less than a factor of 2. These r esults suggest that the models used were insensitive to changes in As speciation (V vs . III). This would not be the case under real soil / environmental conditions where mu ch lower mobility of As(V) from CCW compared to As(III) would result in reduced risk (much greater than a factor of 2) for As. The differences in As transport could be ove restimated in orders of magnitude because As(III) was used instead of As(V).

The results should be modeled using As(V). The Kd values for As(V) and As(III) used for inputs into EPACMTP should be reported. The MINTEQA2 outputs used to calculate Kd values should be reported also.

Dr. Charles Harvey, Massachusetts Institute of Technology Numerical modeling of reactive transport is a challenging area of research. Constructing an accurate reactive transport model for a singl e site is a very ambitious research goal, even with abundant site-specific data. So, of course, calculating reactive transport over a wide variety of possible sites requires extensive simplifications.

The models employed here simplify many geochemical processes, and it is easy to construct a list of biogeochemical processes th at are not included. This list could contain processes that would either reduce or increase retardation. Geochemists who study surface reactions may criticize the choice of metal oxides and the parameterization of their surface characteristics; Geochemists who study organic matter may criticize theDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

7 modeled source of organic matter and its reac tivity; Geochemists who are interested in redox transformations may criticize the l ack of redox transformations, such as As(III)/As(V), and the lack of realistic models of microbial metabolisms. All of these concerns, and many others, are likely valid at some sites, but not others, for some contaminants, but not for others. In other wo rds, our ability to model reactive transport may not be sufficiently developed to co nstruct good models over a wide range of conditions without constraini ng these models with a very large data set of field
measurements.

So, does is this assessment over-reach by employing a reactive transport model? Is constructing an assessment across many sit es a hopeless task? Perhaps, but we know some simple constraints on the spread of contaminants. First, contamination is not transported farther than a conservative tracer. Se cond, the spread of most contaminants is considerably less than for a tracer because they adsorb to the aquifer material, and perhaps precipitate. Thus, it seems to me, the tact taken here, of reducing the effects of reaction to a set of different simple isothe rms, and then using these isotherms to model contaminant immobilization and retardation of transport, may be the only realistic approach to this problem.

Perhaps my biggest concern about the reactive transport model is that that it appears to use parameters that would represent a mixture of the leachate and ambient conditions. For example, appendix D describes how the pH was calculated as “The volume of leachate released from the WMU was mixed with the volume of the aquifer that was likely to be impacted by the plume”. This approach misses an important fact about solute transport in aquifers: the volume of mixed water may be ve ry small; most of the groundwater is either in the plume, or out of the plume. The mixing zone may be very thin. Correctly modeling a reacting plume is very difficult, but the simple approach taken here would likely be more realistic if the conditions within the pl ume were used to calculate the retardation factor.

One simple, but important, aspect of the model should be clarified. With the retardation approach, not only is the residence time of a so lute increased by its retardation factor, but the concentration must also be decreased by the retardation factor. I suspect that the models used in this study correctly reduced the concentration by the retardation factor, but I found the text confusing on this issue. If this simple effect was not included it would be a serious error.

EPA Response EPA recognizes that explicit reactive/geochemical mode ling would be more mechanistic than using linear and nonlinear partitioning coefficients; however, su ch modeling is not a practical approach for a nationwide analysis because (1) the data collecti on effort necessary to populate such a model on a nationwide, site-based level would be prohibitively expensive; and (2 ) assuming data were available to populate a reactive model, the level of complexity and time necessary to run such a model would also be impractical. For a broad, nationwide analysis, the use of Kd is a practical and necessary simplification. The use of Kd as a surrogate for dilution/precip itation processes is a widely known and accepted method in the scientific literature and the groundwater mode ling community, provided the values of Kd used are appropriate.

Nevertheless, we have added discussion to the risk assess ment to clarify issues raised by the commenters. Appendix D of the revised risk assessment displays selected percentiles of the Kd estimates. These estimates were derived from the isotherm sampling performed by EPACMTP and used in the modelingDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

8 (including effective Kds for the unsaturated zone). A complete listing of all Kd values available in the MINTEQA2 isotherms used in these analyses would not be practicable.

We considered comparing the modeled Kd values to available estimates in the published literature, but did not for two reasons. First, there are many indivi dual values in the Kd isotherms (depending on metal concentrations, and master variables such as pH, orga nic carbon, and iron oxide concentrations). Second, there is limited information generally available a bout the conditions under whic h published values were measured. This comparison would be very time cons uming and subject to many uncertainties because of the complexity of geochemical conditions that impact Kd values in both field and in laboratory studies.

With respect to the use of goethite versus the use of hydrous ferric oxide, EPA had discussions with Dr. David Dzombak and Dr. Samir Mathur (developer of the goethite database). In these discussions, the group discussed the sorbent question extensively, and EPA chose to use goethite rather than hydrous ferric oxide as a best estimate that would not underestimate risk. As pointed out by Dr. Harvey in his review, the question of sorbents is “likely valid at some sites, but not others,” and “using these isotherms to model contaminant immobilization and retardation of transport, may be the only realistic approach to this problem.” EPA agrees that this was a necessary assumption and has added discussion of this uncertainty to Section 4.4.2.

The approach adopted in the risk asse ssment to determine the value of pH in the aquifer (used to select Kd) and the subsequent calculation of the retardation factor assumed that, after entering the aquifer, the leachate plume would thoroughly mix with the ambient, uncontaminated groundwater. The commenter suggests that the mixing zone is largely at the periphe ry of the groundwater plume, which is consistent with the general conceptual model used in this risk assessment (essentially uniform flow with recharge). EPA adopted the assumption of full mixing as a more conservative approach, because for most metals, sorption/precipitation tends to increase (i.e., Kd goes up) with higher pH, which is characteristic of much CCW leachate (i.e., assuming full mixing lowers the groundwater pH and thus decreases sorption).

For this response, we conducted a brief analysis to compare modeling results based on the assumption that little or no mixing occurs between the entering leachate and the groundwater versus results based on the original assumption that the leachate fully mix es with the groundwater. The comparison aims to assess the consequences of adopting the conservative, fully mixed approach. To bound the problem, the analysis compares two landfill Monte Carlo simulations for coal ash waste containing As(III) and coal ash waste containing As(V): one applies the risk assessment as sumption of using a fully mixed aquifer pH for Kd selection and the other applies an assumption that no mixing occurs in the aquifer by using the leachate pH as the governing pH for Kd selection. These two sp ecies were selected specifica lly for their respective sorption isotherm behavior; Kds derived from As(III) is otherms tend to decrease as pH increases (which is typical of most metal species examined in th e risk assessment), whereas Kds derived from As(V) isotherms tend to increase with increasing pH.

Percentiles of peak receptor well concentration from both simulations were selected and compared by calculating the percent change as follows:



### 100 C C C

Mix Full Mix Full Mix No 

### × −

= Change %

where

C No Mix = Simulated peak receptor well concentra tion for a select percentile based on a no mixing assumption (mg/L)DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

9 C Full Mix = Simulated peak receptor well concentrati on for a select percentile based on a fully mixed assumption (mg/L)

Table 1 shows the percent change in peak receptor we ll concentrations assuming no mixing occurs for a range of percentiles.

Table 1. Change in Peak Receptor Well Con centrations for Ash Disposed in Landfills Assuming Leachate Does Not Mix in Aquifer

Percentile of Peak

Concentration Percent Change in Peak Concentration

As(III) As(V)

10 0.00% 0.00%

20 0.00% 0.00%

30 0.91% 0.00%

40 0.25% 0.00%

50 0.31% 2.28%

60 0.00% 15.57%

70 0.23% 57.97%

80 0.00% 18.31%

90 0.00% 11.75%

The results for the two arsenic speci es indicate that As(V) has a sensitivit y to pH that leads to increased receptor well concentrations under the no mixing assump tion (i.e., when the leachate pH is used to determine Kd in the saturated zone) relative to th e well mixed assumption used in the risk assessment. These results suggest that a change in the leachate mixing approach could lead to a change in risk for some constituents. A discussion about the uncertainty asso ciated with the approach for estimating pH has been added to the model uncertainty section (4.4.2) of the revised risk assessment, including the above table as Table 4-28.

Dr. Basta points out that literature on arsenic V often shows it to be orders of magnitude less soluble than arsenic III, and that EPA’s results differing by a factor of two is inconsistent with these findings. To deal with the problem of speciation, EPA now reports results assuming 100% arsenic III and 100% arsenic V. While the original comparison Dr. Basta refers to was made at the 90th percentile, EPA notes that the 50th percentile results often do differ by the orders of magnitude that Dr. Bast a suggests. For instance, although the 90th percentile results for conventional CCW in unlined landfills differ by a factor of two (4×10 -4 for arsenic III versus 2×10 -4 for arsenic V), the 50th percentile results differ by several orders of magnitude (6×10 -6 for arsenic III versus 6×10 -10 for arsenic V). The lack of sensitivity to speciation for the high-end risk results is most likely because the mode l runs with higher risk typically involved shorter distance to receptors. Thus, one would expect less ti me and distance for adsorption and precipitation to occur. Updated discussion of the uncertainties with ar senic and selenium speciation is reported in Section 4.4.2 of the revised risk assessment.

Finally, with respect to Dr. Harvey’s comment, th e contaminant transport fo rmulations in both the unsaturated and saturated modules of EPACMTP correc tly include a sorption term that represents the transfer of dissolved contaminants from the liquid ph ase to the solid phase by way of a retardation factor or an isotherm describing the change in sorbed concen tration with respect to dissolved concentration. TheDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

10 current risk assessment report (Appendix D, Section D.1, paragraph 1, 5th sentence) has been modified by adding the following text to better explain this: “The retardation factor (R) accounts for the effects of equilibrium sorption of dissolved constituents onto the solid phase, removing them from solution and reducing the available mass in the dissolved phase.”


### 3.3 Site-Specific Analysis and Co mparisons with Measured Data

Comment Summary Three reviewers (Dr. Basta, Dr. Harvey, and Dr. Vorhees) submitted comments that fall into this subcategory. Dr. Basta and Dr. Vorhees recommended conducting more detailed, site-specific analysis to increase c onfidence in the risk assessment results. They expressed concern that the results are difficult to evaluate given the significant variability and uncertainty associated with the national scope of the analysis.

Dr. Nicholas Basta, Ohio State University EPA is to be commended on the great e ffort expended to conduct a nationwide assessment of risks posed by CCW disposal pr actices. However, is it possible to model human and ecological risks across the wide range of scenarios found nationwide at different CCW facilities? Confidence in such an all-encompassing model can only be achieved with (1) verification or “ground-truthi ng.” There is a wealth of field leachate and impact data from CCW disposal facilities. Because of the great uncertainty in the model, the model should be validated on a case-by-case basis with the CCW field data to gain acceptance for the EPA CCW risk assess ment model. Perhaps a subset of CCW facilities can be selected for verification of the risk assessment model output. This validation of model results must be perf ormed before any decisions on CCW waste management are required of specific CCW facilities.

Dr. Charles Harvey, Massachusetts Institute of Technology The model described in this report predicts th e distribution of contaminant concentrations downstream of WMUs. The obvious next step is to test the model by comparing it to measurements. This is easily accomplished by simply (2) collecting and analyzing some water samples, then testing the hypothesis that the measure d values come from the same distribution predicted by the model. This is the scientific method.

Dr. Donna Vorhees, The Science Collaborative Need for More Site-Specific Analysis. The primary challenge in conducting this review was to interpret results of this ambitious , national-scale PRA in light of the myriad sources of uncertainty associated with the exposure scenarios and risk models. The national scale of this assessment introduces co mplexity and uncertainty that perhaps could be reduced if EPA examined a represen tative subset of WMUs in more detail to better understand the technical strengths and weaknesses of its nationwide approach. One starting point might be those WMUs that are represented in existing surveys of WMUs (i.e., EPRI, 1997 and [U.S.] DOE, 2006) a nd in EPA’s CCW constituent database. For example, I am aware of one site referenced in Figure 4-5 that was the subject of a relatively recent risk assessment that was subm itted to the Massachusetts Department of Environmental Protection. Public commenta ry submitted from industry and government (State of Wisconsin) representatives noted th e availability of additional data, and these leads should be explored to determine if hi gh quality data are available beyond what is already included in the risk assessment. EPA does make reference to WMUs involved in damage cases, but these WMUs are not necessarily representative of all WMUs. EPA’sDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

11 risk assessment embodies an enormous amount of work and creative, thoughtful modeling. But, at a minimum, EPA need s to provide a compelling reason for not attempting some site-specific analyses, wh ich is the most effective way to reduce uncertainty in the PRA.

EPA reports that it has data for antimony, coba lt, mercury, and thallium from, at most, “a few” sites and with “only a few measurements” (page 4-32). EPA should explain why this is the case (e.g., elevated detection li mits, not included on analyte lists at surface impoundment WMUs). Screening results in Table 2-3 suggest that it would be worthwhile to resolve whether these constituents are present and at what concentrations.

EPA Response EPA agrees that confidence in environmental trans port model results increases significantly when model predictions can be compared favorably with measured results such as actual field data. However, sitespecific modeling involves extensive data collecti on and more detailed modeling (representing sitespecific conditions and processes). Instead, EPA has designed the models to support national probabilistic, Monte Carlo analysis representing a wide range of conditions that could occur over time.

The Agency has also found that CCW contaminants of concern in the damage cases agree with those exceeding the risk criteria in this analysis, building confidence that the risk assessment captures national conditions. Table 2 compares the results from the 2007 draft risk assessment with the damage cases reported in the Coal Combustion Waste Damage Case Assessments (U.S. EPA, 2007) for the groundwater pathway.



### Table 2. Modeled 90th Percentile Risk Resu lts vs. Reported Groundwater Exceedences

Damage Cases b

Constituent 2007 Risk Assessment a Human Health Effects c Cosmetic/ Aesthetic Effects d State Standard e Consistent Results as of 2007 f

Aluminum – – T – Yes Antimony T – – – No
Arsenic T T
– T Yes
Barium – – – – Yes Beryllium Screened – – – Yes

Boron T T
– T Yes
Cadmium T T – – Yes
Chloride Not Screened – T T N/A
Chromium RAF T – – Uncertain
Cobalt T – – – No Copper Screened – – – Yes

Fluoride RAF – – T Uncertain
Iron Not Screened – T T N/A
Lead T T – – Yes
Lithium Not Screened –

– T N/ADRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

12 Damage Cases b

Constituent 2007 Risk Assessment a Human Health Effects c Cosmetic/ Aesthetic Effects d State Standard e Consistent Results as of 2007 f

Manganese RAF T T T Uncertain Molybdenum T T – – Yes
Nickel RAF T
– T Uncertain
Nitrate/Nitrite T – – – No
Selenium T T
– T Yes
Silver Screened – – – Yes
Sulfate Not Screened – T T N/A
Thallium T – – – No
Vanadium RAF – – – Uncertain
Zinc Screened – T – Yes

a Not Screened = Constituent was not considered due to lack of health-based benchmarks.

Screened = Constituent showed no risk potential in the screening assessment. RAF = Constituent showed risk potential in the screening assessment, and was analyzed with risk attenuation factors. T = Constituent underwent full probabilistic modeling and was shown to pose a risk to human health in the landfill scenario, th e surface impoundment scenario, or both. – = Constituent underwent full probabilistic modeling and was not shown to pose a risk to human health

b T = At least one proven damage case show ed an exceedence of this constituent. – = No proven damage cases have yet shown an exceedence of this constituent.

c T = Exceedences of primary maxiumum contaminant levels (MCLs) or other healthbased numbers published by EPA.

d T = Exceedences of secondary MCLs, which would not result in harm to human health.

e T = Exceedences of a relevant state standard.

f Yes = Results of risk assessment and damage cases either both indicated a risk to human health or both indicated no risk to human health. No = The risk assessment indicated risks where none have yet been found in a proven damage case. Uncertain = It is possible that the results were consistent, but due to lack of probabilistic modeling, no definitive conclusion can be made. N/A = Lithium was not examined at any stage in the 2007 ri sk assessment, so it was not possible to draw any conclusions as to consistency.

EPA also acknowledges that the statement on page 4-32 is mistaken with respect to cobalt. This has been corrected in the revised risk assessment.

The first category of cons tituents are those for which the risk assessment and the damage cases agree. Some constituents in this group showed risk in the m odel and also appeared in the damage cases (arsenic, boron, cadmium, lead, molybdenum, and selenium). Othe r constituents in this group did not show the potential for risks in excess of the risk criteria from th e risk assessment and did not appear in the damage case literature (aluminum, barium, beryllium, copper, silver, and zinc).DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

13 The second category of constituents are those for whic h the risk assessment and the damage cases did not agree. Four modeled constituents (antimony, cobalt, th allium, and nitrate/nitrite) showed risk at the 90th percentile but no damage cases had been proven as of 2007. This could indicate that (1) the risk assessment was conservative for th ese constituents, (2) not enough time has passed to see the remaining constituents appear in damage cas es, (3) corrective action was taken when the first constituent(s) was observed, so further constituents that would have appeared at the same site were never seen, or (4) these constituents are not tested for as frequently as the constituents found in the proven damage cases.

The third category of constituents are those that were not screened out and were analyzed using risk attenuation factors (chromium, fluoride, manganese, ni ckel, and vanadium). Because all that is known is that these constituents have the potential to pose a risk to human health, th ey cannot currently be compared to the damage case results.

The final category of constituents are those that were not evaluated at either the screening or modeling stages because no health-based values were available for comparison. These four constituents (chlorine, iron, lithium, and sulfate) appeared in damage case s due to exceedences of aesthetic or state levels, not due to a known risk to human health.

Table 3 compares the results from the 2007 draft risk assessment with the dama ge cases reported in U.S. EPA (2007) for the fish consumption pathway. The only fish consumption advisories documented in CCW damage cases are for selenium. This is consiste nt with the risk assessment for selenium. The two constituents that do not pose a risk in the risk assessment (cadmium and thallium) were also not part of any fish consumption advisories in the damage cases. The one inconsistency is arsenic, for which the risk assessment shows a cancer risk of 1 in 50,000, which slightly exceeds the risk criteria of 1 in 100,000. However, no arsenic fish consumption advisories exis t at proven damage case sites. This inconsistency could indicate that (1) the risk asse ssment was conservative with resp ect to arsenic, (2) not enough time has passed to see arsenic appear in fish advisories at these sites, or (3) the arsenic exceedences have not been detected in random fish tissue samples thus far.



### Table 3. Modeled 90th Percentile Risk Results vs. Reported Fish Consumption Exceedences

Constituent 2007 Risk Assessment a Damage Cases b Consistent Results as of 2007 c

Arsenic T – No
Cadmium – – Yes
Selenium T T Yes
Thallium – – Yes

a T = Constituent underwent full probabilistic modeling and was shown to pose a risk to human health in the landfill scenario, th e surface impoundment scenario, or both.

b T = At least one proven damage case showed a fish consumption advisory for this constituent. – = No proven damage cases have yet shown a fish consumption advisory for this constituent.

c Yes = Results of risk assessment and damage cases either both indicated a risk to human health or both indicated no risk to human health. No = The risk assessment indicated risks where none have yet been found in a proven damage

case.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

14 The above discussion and tables have been incorporated into Section 4.1.5 (as Tables 4-19 and 4-20).



### 3.4 Well-Related Issues

Comment Summary Comments relating to the number of wells cont aminated, the realistic risk of exposure, well placement within the plume, and other well-related issues were received from Dr. Harvey, Dr. Rowe, and Dr. Vorhees. Dr. Ha rvey’s comments focused on the assumption that the wells used in this assessment are c ontaminated (i.e., located within the plume), even if the well location used reflects a deeper well that may be screened in an uncontaminated aquifer; the manner in wh ich the assessment handles uncontaminated wells, plume characteristics, groundwater–s urface water interactions and vertical contaminant concentration across a screened inte rval in a aquifer; and the values used for plume dispersivity. Drs. Rowe and Dr. Vorhees commented on this assessment’s use of water well distances from municipal solid waste (MSW) landfills and EPA’s belief that these distances would be protective for CCW waste.

Dr. Charles Harvey, Massachusetts Institute of Technology At the heart of this assessment is a Monte Carlo simulation that first generates realizations of contaminant breakthrough curv es at a well that has been contaminated by a WMU. These realizations provide a set of exposures that are then used to determine a distribution of health risks. This approach does not attempt to calculate a probabilistic description of the number of wells that ar e contaminated. Rather, this Monte Carlo approach characterizes the distribution of breakthrough curves assuming that a well is contaminated.

This approach has merits, but it also pre sents puzzling conceptual problems stemming from the fact that the calculated health risk s are independent of the probability that the contaminated well actually exists. The follo wing example illustrates the difficulty. In comment II.B. below, I will discuss how, wherev er there is a river or stream between the source and the well, groundwater will likely discharge into the river, preventing the plume from reaching more distant wells. Consequentially, wells far from the source cannot be contaminated and health risks are reduced. However, in th e framework used in this report, the calculated health risks w ould actually increase for a shorter plume – reducing the plume length forces the Monte-Carlo model to select a well closer to the source where concentrations are higher and arrival times are quicker. Since every realization of the groundwater exposure pa thway must generate a well position, short plumes will actually have worse adverse hea lth effects because the wells generated by the Monte Carlo study will be closer to the s ource where contamination concentrations are worse. Thus, a reduction in the probability of exposure does not necessarily reduce the probability of adverse health outcomes in this model. This problem also exists when defining plume width (see II.B below). It is a fundamental problem of the framework of this assessment that extends beyond these exam ples. The difficulty arises because this assessment does not attempt to characterize the number of wells contaminated, and hence does not consider the size of the population th at would suffer adverse health effects.

This report would benefit from a candid discussion of these issues – a straightforward explanation that the calculated health dangers are conditioned on a well being contaminated, but that in fact many wells, or few wells, could be contaminated. AfterDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

15 reading the public comments, I believe that this discussion would be valuable --it appears that the authors of these public comment s may not have understood this issue.

Public understanding of the dangers of c ontaminated groundwater down-gradient of potential sources of contamination has grown over the last decades so that it is unusual to site new wells, or to leave existing wells, in the flow-paths down-gradient of obvious contaminant sources such as a waste site. B ecause studies such as this have documented the serious health dangers of contaminated groundwater, and because severe contamination has often been found down-gr adient of waste sites, people and local government are aware of the dangers and have taken action to avoid the use of contaminated water. Although the risk to people drinking contaminated groundwater certainly still exists, it has receded. Again, this report would be improved by a discussion of the issue of contaminated wells actually being used for water supply.

The report nicely integrates the few available national data sets of hydrogeolgic parameters. My comments largely concern th e interaction of the choice of well position with the solute transport model. The Monte Carlo simulation places a well at random (uniformly distributed) angular offset from the plume center line, at a depth generated from a uniform distribution of depth between the aquifer top and bottom, and at a distance from the source generated from a dist ribution of distances measured in 1988 for WMUs. This framework is basically reasona ble. However, the implementation is not fully explained and there may be some opportunities for improvement.

(A) Definition of the Plume Width

The definition of the plume size is very impor tant for this assessment, but it is not clear how the margin of the plume is defined. Contaminant concentrations typically fall-off gradually away from the center of a plume. Therefore, the choice of the width of the plume is somewhat arbitrary – it depends on the minimum concentration used to define the extent. The margin could be defined as th e outermost flow path that originates from the waste site. (i.e. the plume that would be simulated with zero dispersion) However, this definition would exclude contamination that disperses out of this flow path. So, a better definition of the plume extent would use a minimum concentration – but what concentration should be used?

This problem of plume definition plays direc tly into the conceptual problem described above in comment I.A. If a low concentration is chosen to define the plume margin, then the well may be placed in a much larger area of low concentrations, and hence the calculated health risks will be small. If the plume margin is defined by a high concentration, then the well will be placed in a high concentration region, and the calculated health risks will be worse. In either case, the simulated plume is the same – but defining the margins with a low concentration su ch that the plume is larger will lower the health risks.

(B) Transverse Dispersion

In a steady-state flow model, such as used in this study, the lateral spread of the plume is largely controlled by the transverse disper sivity. Consequentially, the transverse dispersivity is the key aquifer parameter for determining both the width of the plume and the dilution of contaminants. Recent field and laboratory experiments show that there is much less spreading in the transverse directi on than in the direction of flow, and henceDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

16 the actual values of transverse dispersivity are very small, so small that is difficult to simulate solute transport accurately with many numerical techniques that attempt to model dispersion. I find no mention in the report of the values used for transverse dispersivity. This report should: (1) Give the values used in the transport simulations, (2) Describe how this value was chosen, and (3) Describe the sensitivity of the results to the chosen dispersivity value.

(C) Depth of the Plume

Vertical spreading is much like transverse spreading – little spreading occurs in the vertical direction. Realistic dispersivities are so small that they are difficult to include in simulation models. So, again, one wonders what values were used, and how the simulation was performed.

However, the effects of vertical spread ing on contamination of wells differs from horizontal spreading because wells are often scr eened over large vertical increments so wells often extract water that contains a mix of concentrations over a vertical extent of the contaminated aquifer. Wells can be screened over 10m, and, depending on the pumping rate, can draw water from the full thic kness of the aquifer. Thus, the vertical concentrations might be bette r approximated by averaging concentrations vertically.

Finally, geologic formations are often layere d with deep confined aquifers separated from shallow unconfined aquifers by low permeability layers. Wells are often placed in the deeper aquifer, but local contamination is ofte n restricted to the upper aquifer. Some of the data used to construct the distribution of well locations may reflect wells that tap deeper aquifers that will remain uncontaminated by WMUs.

(D) Rivers May Limit the Distance from the Source

I suspect that many waste sites are near rivers because power plants are often along rivers. Except in arid regions, rivers are typically gaining, meaning that groundwater discharges into the river. Indeed the mean distance to surface water calculated in this report is less than the mean distance to a well. Consequentially, contaminant plumes discharge into rivers, and hence the maximum length of a plume is limited by the distance to a river. It is not clear from figure 3-7 that plumes are constrained by the distance to surface water, although these distances are used for the calculation of risks to fishermen and ecological risks.

Dr. Kerry Rowe, Queens University In the absence of an actual survey of wa ter well usage around existing CCW facilities, I am not aware of a better source of data than that available for MSW landfills. EPA recognized this as “one of the larger uncer tainties in this analysis” (RA, pg. 4-37) however the statement that “EPA believes th at the MSW well distance distribution used is protective for CCW landfills and surface im poundments” (RA, pg. 4-37) appears to be nothing more than that – a belief – with no ju stification being given for this article of faith. Also, as noted in response to Question #5 , the analysis extends for 10,000 years but no consideration appears to have been gi ven to likelihood of increased demand for the use of groundwater in the next 30-100 years (let alone the next 10,000 years) which presumably means more wells and makes the pr otection of groundwater likely even more important for the future than it is now. It would have been good to have examined the implications of wells being closer than th at indicated by the MSW survey to assess howDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

17 sensitive the RA is to that assumption (one can infer that it will be important – but not how important in terms of risk).

Dr. Donna Vorhees, The Science Collaborative Distribution for Distance to Drinking Water Wells. This distribution is the subject of one of our charge questions:

• Charge Question #11: The CCW risk assess ment used a distribution of well distances

measured at municipal solid waste (MSW) la ndfills. Is this distribution appropriate forCCW disposal facilities? Are there better or more representative sources of welldistance data for CCW disposal facilities that should be used instead?

It might be appropriate, but it is hard to sa y in the absence of any real measurements from WMU locations. One public comment on this distribution refers to its bias, which EPA describes as conservatism. I agree that such bias should be avoide d in conducting PRA. EPA does not explain why it could not obtain measurements from at least a representative sample of WMUs.

EPA explains that:

“...the MSW well distance distribution is a conservative representation of actual well distances at CCW disposal sites. However, data on residential well distances from CCW landfills or surface impoundments will be needed to verify this hypothesis” (Appendix C, Section C.2).

This sentence seems to suggest that EPA w ill perform additional analyses at some point and possibly update the risk assessment. This would indeed be a useful step, and EPA’s intentions should be clarified.

EPA assumes that the drinking water pathway exists at all WMU locations, using the distance between municipal landfills and dri nking water wells as surrogate measures for the distance between WMUs and drinking wa ter wells. EPA appropriately cites the uncertainty associated with this surrogate in formation but does not explain the specific impediments to obtaining data specific to at least a representative subset of WMUs. Because everyone must drink water, presumably everyone served by these hypothetical wells would be exposed, although EPA correctly notes that people are likely to get some or all of their water from other sources. Th is exposure pathway might not occur near every WMU and assuming so contributes to the uncertainty in the risk output distribution of risk for this pathway.

EPA Response EPA agrees with Dr. Harvey’s comments that in case s where wells are located beyond an intersecting surface water body, the draft risk assessment may not account for interactions between surface water and ground water. Examining the input database, EPA notes th at approximately two-thirds (69 percent) of the Monte Carlo runs contained such an intersecting surface waterbody. However, the WMUs with closer receptor wells exhibited higher risks on average. Thus , the 90th percentile results are not likely to be significantly affected.

The risk assessment now includes a more detailed expl anation of how mass is “transferred” from the groundwater to the intersecting surface water body, and more clearly explains that the groundwater plume is assumed not to be affected by surface water/ groundwater interactions in Section 3.6.3:DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

18 “The waterbody was constrained to lie acr oss the contaminant plume centerline and perpendicular to the groundwater flow dir ection. The waterbody is assumed to fully penetrate the aquifer thickness.”

and in Section 3.7:

“The primary simplifying assumptions in EPACMTP are as follows: (1) the groundwater–surface water interface is assume d to be perpendicular to the regional groundwater flow direction; (2) the interface is infinite in its lateral extent so as to intercept the entire width of the dissolved contaminant plume; and (3) the intercepting surface water body fully penetrates the saturate d region of the subsurface. Therefore, all of the mass in the contaminated groundwater is available to be transferred to the surface water model. If stream flow is greater than th e available groundwater flow, then all of the mass available in the groundwater is assumed to be transferred to the surface waterbody. It is important to note that while a mass tran sfer is assumed to take place between the two systems, mass is not actually removed from the groundwater—it is still available to be observed at a receptor well placed beyond th e groundwater-surface water interface.”

The comments regarding the placement of receptor well s in the Monte Carlo framework are the result of a fundamental misunderstanding about the constraint of the groundwater receptor location as being “within the contaminant plume” as described in Section 3. 4.3 of the 2007 risk assessment document. This constraint was not well explained. Thus, the explanati on beginning in paragraph 4 of Section 3.4.3 of the 2007 document (now Section 3.6.3 of the revised document) has been rewritten to more clearly describe the methodology used to determine the constrained la teral position of the receptor well. With respect to Dr. Harvey’s comment on the assumption that deeper wells were contaminated, EPA has consistently adopted a practice of erring on the side of protecting human health and environmental quality with respect to modeling assumptions. Consistent with this prac tice, EPA will continue to focus on wells screened within vulnerable, surficial aquifers (i.e ., top 10 meters of the saturated zone).

In the comment on vertical concentration gradients, Dr. Harvey highlights the possibility that modeled receptor well concentrations may over- or underpredict th e actual level of exposure. This possibility exists because simulated concentrations are point observa tions, not concentrations from simulated pumping wells that may capture both contaminated and pristin e groundwater, depending on factors identified in Dr. Harvey’s comment. Nevertheless, EPA opted to u se the current approach, which may conservatively overestimate the receptor well concentrations.

Dr. Harvey’s comment on dispersivity partially highl ights the issue of transparency. The documentation for this risk assessment did not include comprehens ive tables detailing modeling input parameters, their values or distributional characteristics, and the sour ce of the data used; however, these values are, in many cases, provided in the EPACMTP Background a nd Parameters/Data Background documents (U.S. EPA, 2003a,b) and in general, we did not duplicate that publically available material in the risk assessment document, but do refer the reader to it wher e appropriate. For example, Section 3.4.2.1 says

“The methodology, assumptions, and data used to determine infiltration rates for each CCW liner scenario were consistent with the approach used in EPA’s Industrial D guidance, as described in Section 4.3 and Appendix A of the EPACMTP Parameter/Data Background Document (U.S. EPA, 2003a) and Section 4.2.2 of the Industrial Waste Evaluation Model (IWEM) Technical Background Document (U.S. EPA, 2002b).”

In addition, EPA’s comprehensive input data files can be made available in Microsoft Access format upon request. As to including a sensitivity analysis of the results with respect to values used for the dispersionDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

19 tensor, it is EPA’s opinion that the need for an a dditional sensitivity analysis will be mitigated by improvements in the document with respect to a more detailed description of the “in the plume” receptor well constraint.

In response to Dr. Rowe’s and Dr. Vorhees comments on the use of MSW landfill water well distances, the last sentence of Section 4.4.3.3, Paragraph 2, has been revised to include a reference to discussions in Section C.2 of Appendix C of the risk assessment, which provides more detail on the subject. It is particularly important to note that if distances from co al waste landfills were available, EPA would have used them in the risk assessment. However, given limited data, the distance from municipal landfills was chosen as the best approximation.



### 4.0 SOURCE MODELING The comments relating to source modeling are divided into 12 subcategories:
1. The use of leachate concentrations based on t oxicity characteristic leaching procedure (TCLP) data and the availability of more representa tive results from EPA’s research with Vanderbilt University 2. The use of 2006 Department of Energy (DOE) facility data and liner scenarios 3. Alternative waste disposal/codisposal scenarios 4. Updates on liner leakage assumptions

5. Treatment of nondetect CCW constituent values 6. The effect of waste compaction on infiltration and leaching 7. Hydrologic Evaluation of Landfill Performance (H ELP) model infiltration rates and landfill cover assumptions 8. Scenarios where the water table is above the disposal unit 9. The impact of increased mercury capture under future air emission control regulations 10. The use of porewater data for su rface impoundment source concentrations 11. Modeling surface impoundmen ts only before closure 12. Modeling landfills only after closure.



### 4.1 Use of TCLP Data and the Ava ilability of the Vanderbilt Study

Comment Summary This subcategory addresses comments from Dr. Basta, Dr. Hopkins, Dr. Rowe, and Dr. Vorhees. These reviewers expr essed concern about the hierarchical process of choosing data from different leaching methods and the predominant use of TCLP data to derive source concentrations for the risk assessmen t, because the TCLP laboratory method is intended to mimic conditions in municipal landfills rather than CCW monofills. In the presence of organic acids associated with m unicipal waste degradation, the mobility of metals can be significantly impacted by the more acidic conditions and complexing organic acid molecules. Because organic acids are not present in CCW, leachate concentration estimates based on TCLP results may differ significantly from leachate concentrations under actual CCW field conditions.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

20 Dr. Nicholas Basta, Ohio State University In the Draft (p. 4-32), the importance of leachate data is acknowledged in the quote “...waste leachate concentration is arguably the most important data set in terms of driving the risk assessment r esults.” Therefore, additional information that improves the waste leachate data set would be expected to significantly reduce key uncertainties. Selection of appropriate analyte data for la ndfills in the CCW constituent database (p. A4) is a hierarchical process of data from different leaching methods. The hierarchy is described as (a) landfill leachate or porewater data given greater preference than TCLP or SPLP, and (b) selection of the highest concen tration between SPLP and TCLP. The Draft recognizes the quality of the data is much greater for porewater than TCLP. Then, the Draft uses mostly TCLP data as input.

The TCLP method used to generate most of the input leachate data in the Draft is improper for determining leachate concentr ation from CCW. Improved leachate methods, designed for CCW, should be used to gene rate data in the Risk Assessment.

TCLP was developed to characterize waste beha vior in municipal landfills. It is not a predictive method. TCLP extracts are acidic (pH < 5), buffered with acetic acid, and designed to simulate decaying organic waste (i.e., food, etc.). However, conditions are very different in CCW landfills where organic acids from municipal solid waste are absent and conditions are more likely to be neutral to alkaline than acidic. TCLP is a gross extraction to classify disposal of waste materials in either hazardous or nonhazardous landfills. TCLP was not designed to model specific physiochemical processes such as leachate from CCW landfills. Therefore, TCLP data should not have been used to perform risk assessment calcula tions of CCW. The following is stated in the Draft (p. 4-32) recognizing the limitations of the TCLP method: “The available landfill data are largely derived from the TCLP, a la boratory test designed to estimate leachate concentrations in municipal solid waste (MSW ) landfills.” The use of TCLP data in the risk assessment is justified as follows:

“The TCLP has been shown to both over and under predict leachate concentrations for other waste disposal scenarios, so the use of the TCLP data to represent CCW leachate is another source of uncertainty. However, as noted below, this does not appear to be a significant source of uncertainty for this analysis.”

To support this contention, limited results of leachate data from CCW material from a study conducted at Vanderbilt University (Vanderbilt, 2006) were compared with extraction data used in the CCW Risk Assessm ent for As and Se (Fig. 4-5). Overlap of the ranges of As concentration reported in Vanderbilt (2006) and the Draft CCW for 5 of the 6 CCW facilities is used to conclude:

“Although the Vanderbilt Study does not cover all of the metals addressed in the CCW analysis, its general agreement with the CCW arsenic and selenium levels does help allay concerns that the TCLP CCW leachate values u sed in the analysis markedly overestimate or underestimate the concentration’ s actual CCW leachate” (4-34).

This is not a sufficient justification for use of TCLP data as acceptable data because:

(a) A comparison of results from six CCW facilities is inadequate. To accept TCLP data as an acceptable data, a comparis on should be made between TCLP data and actual leachate for all CCW materials.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

21 (b) The comparison between TCLP and actu al leachate data would have to be extended to include all the input da ta used in the risk assessment.

Accurate input data is critical to determine risk for a specific site. For example, data present in the Draft from Vanderbilt (2006) for th e St. Clair facility of 0.001 mg/L for As is 60 times lower than ca. 0.06 mg/L used in the Draft Risk Assessment (Fig. 4-5).

Recently, research has improved the scientific understanding of the generation of leachate from CCW, in particular for mercury, arsenic, and selenium (USEPA, 2006) and development of new CCW-specific leac hing methods (USEPA, 2006e; Kosson et al., 2002). Improved leachate methods, designed for CCW, should be used to generate data in the Risk Assessment.

Dr. William Hopkins, Virginia Polytechnic Institute and State University It is widely recognized that TCLP data ar e not adequate for characterizing how coal combustion wastes will behave in a field se tting. There are a number of serious problems with using this test to predict long-term be havior of these wastes, some of which were highlighted in the National Academy of Sciences 2006 report (pages 143-152). Unfortunately, TCLP data may simply be the b est available for the time being, but when better alternatives are developed (there ar e ongoing, active efforts to do so) the EPA should determine whether the outcome of those tests suggest a re-evaluation of this risk
assessment.
Dr. Kerry Rowe, Queens University The use of landfill data largely based on TC LP is potentially subject to criticism. An attempt is made to address this issue (e.g. RA , pgs. 4-32 to 4-34). At least for arsenic and selenium a good case is made that the paramete rs used in the RA are reasonable for arsenic and or potentially unconservative for sel enium. However it is not clear from the documentation how one can validly distinguish between the leachate from conventional CCW and CCWs co-disposed with coal refuse if data is based on the TCLP. One would anticipate that potentially higher concentrations of metals in the more acidic leachate associated with co-disposal than for conventional CCW. There maybe data in the Vanderbilt study that provides insight regardi ng this matter however it is not addressed in the current RA.

The use of landfill data largely based on TCLP is potentially subject to criticism, especially for the CCW given the pH used in the TCLP test. However the report does address this concern (pgs. 4-32 to 4-34) and at least for arsenic and selenium a good case is made that the parameters used in the RA are reasonable for arsenic and or potentially unconservative for selenium.

If more work were to be done today there is at least the data from the Vanderbilt study however it would appear that if more work is to be done it should also look at the implications of flue gas mercury controls.

Dr. Donna Vorhees, The Science Collaborative Specifically, EPA describes and partially analy zes data that have been collected recently by EPA and DOE (DOE 2006) that reflect changing practices at WMUs. EPA also refers to findings from its study of CCW chemistr y and leaching behavior (EPA 2006b). EPA explains that these data were not included because the risk assessment was performed inDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

22 2003 (page 1-1). EPA acknowledges that inclus ion of these data might result in lower risk estimates. Without reviewing them in detail, it is not possible to say whether inclusion of these data would “significantly improve” the risk assessment, but I see no logic to excluding them. Without inclusion of data that represent more recent conditions such that the database as a whole is re presentative of WMUs used for CCW disposal, EPA’s data set is not of sufficient quality and quantity to support the objectives of the risk assessment. EPA reports that it “is current ly evaluating how to best incorporate and address the results and findings of these studies in future efforts to address CCW management practices” (page 1-1). This stat ement raises another question: how will the results of this risk assessment be used? Are th ey just interim calculations that EPA plans to supplement with future analyses that include these more recent data?

EPA appropriately acknowledges limitations associated with TCLP data and compares it to higher quality data that it collected mo re recently (See Figure 4-5). EPA argues that the data sets are similar but this figure suggests ot herwise. It would be interesting to see how risk estimates change with the exclusion of TCLP data, assuming this can be done while maintaining a representative dataset.

EPA Response For surface impoundments, only porewater data were used . Therefore, this issue is not relevant for the surface impoundment scenario.

For CCW landfills, EPA agrees that TCLP may not be the most appropriate leachate extraction method for all waste streams in all disposal scenarios. Thus , a hierarchy was used for choosing leachate data for the landfill scenario. In the hierarchy (p. A-5) the following priority was assigned to different data types:

(1) Direct samples of landfill leachate OR direct porewater samples from a landfill (2) Waste extractions with long equilibra tion times AND low liquid-to-solid ratios (3) Waste extractions with long equilibration times OR low liquid-to-solid ratios (4) TCLP OR the synthetic precipitation leaching procedure (SPLP) OR other dilute water waste extractions. Vanderbilt University and EPA researchers develope d a new leachate procedure for landfilled CCW. This new leachate procedure was used in Characterization of Mercury-Enriched Coal Combustion Residues from Electric Utilities Using Enhan ced Sorbents for Mercury Control (U.S. EPA, 2006c). Although the 2006 report was focused on mercury, it also addr essed arsenic and selenium. The risk assessment included a comparison and discussion of all the da ta used in the risk assessment versus the Vanderbilt/EPA data for these three elements. This discu ssion can be seen in Section 4.4.3.1. Specifically, EPA notes that:

“...the TCLP data do appear to encompass th e range of variability in CCW leachate concentrations that have been measured in mo re recent studies. [...] For arsenic, the CCW values bracket the range found in the other studies. Se lenium values also agree fairly well for CCW landfill data...”

After the release of the draft risk assessment, the Vanderbilt/EPA leachate procedure was again used inCharacterization of Coal Combustion Residues from Electric Utilities Using Wet Scrubbers for MultiPollutant Control (U.S. EPA, 2008c). This second EPA report, like the first, addressed additional constituents with mercury. These included alumi num, antimony, arsenic, barium, boron, cadmium, chromium, cobalt, lead, molybdenum, selenium, and th allium. EPA does not intend to conduct a new riskDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

23 assessment based on the eight facilities in this new st udy because such a small number of sites would not be as representative of a nationwide distribution as th e numerous sites used in the draft risk assessment. Instead, EPA directs reviewers to the comparisons in Section 3.3.2 of U.S. EPA (2008c). Figures 46–59 of this section show that the data used in the risk assessment and the data in the Vanderbilt/EPA report are generally in the same range for most sites, and thus would not significantly change the risk results.


### 4.2 2006 DOE Facility Data and Liner Scenarios

Comment Summary Dr. Rowe pointed out that the risk asse ssment results represent outdated conditions, because the DOE (2006) study indicated that the proportion of lined versus unlined landfills had increased since the 1995 EPRI data (EPRI, 1997) that were used in the risk
assessment.
Dr. Kerry Rowe, Queens University Objectives 2 and 3 are generally considered to have been met, particularly in the context of the stated intent (RA, §4.4.1, pg. 4- 27) to “characterize current conditions by estimating risks from actual CCW disposal sites across the country” based on data relating to facilities that existed in 1995 as id entified in the EPRI (1997) report. However, as acknowledged in the RA at various points

(e.g. RA, §4.4.1, pg. 4-28), the proportion of lined landfills appears to have changed significantly between the 1995 data and the subsequent U.S. DOE (2006) study. This does not affect the assessment of the risk asso ciated with the facilities that existed in 1995 but the data does need to be carefully inte rpreted if it is to be used to assess the risk of future facilities and for developing futu re management options (Objective 2). For example the combination of data (“All Units Combined”) to evaluate 90th and 50th percentile risks (e.g. RA Tables 4-5 to 4-10) makes sense in assessing the general risk associated with the 1995 facilities but does not represent the risk associated with facilities constructed post 1995. On the other hand th e separately evaluated risks for “unlined”, “clay-lined” and “composite-lined” units can be considered to be useful to the EPA for the development of management options for future CCW facilities (Objective 2).

The limitations associated with the CCW management data being based on the EPRI (1997) sites is acknowledged in the RA (e.g . pg. 4-28). However, this could have warranted a little greater discussion to acknowle dge the issues raised above in response to Question #1.

As discussed in response to Question #1 (a nd repeated here for completeness), the proportion of lined landfills appears to have changed significantly between the 1995 data and the subsequent U.S. DOE (2006) study. This does not affect the assessment of the risk associated with the facilities that exis ted in 1995 but the data does need to be carefully interpreted if it is to be used to assess the risk of future facilities for developing future management options.

The combination of data (“All Units Combin ed”) to evaluate 90th and 50th percentile risks (e.g. RA Tables 4-5 to 4-10) makes sen se in assessing the general risk associated with the 1995 facilities but does not repr esent the risk associated with facilities constructed post 1995. On the other hand th e separately evaluated risks for “unlined”, “clay-lined” and “composite-lined” units can be considered to be useful to the EPA for the development of management options for future CCW facilities.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

24 While the information in RA Tables 4-5 to 4-10 for “unlined”, “clay-lined” and “composite-lined” units is useful for the deve lopment of management options for future CCW facilities, it only considers the specific liner conditions at the specific sites where those conditions actually exist (e.g. the evalua tion of risks for “clay-lined” is only done for sites that are clay-lined). This is appr opriate for evaluating the risk associated with these facilities, however even better insight regarding the effect of the three liner conditions (“unlined”, “clay-lined” and “compos ite-lined” ) could have been obtained by examining all three liner possibilities for a ll sites used in the Monte Carlo study (irrespective of what is currently used) thereby giving an assessment of the risk associated with the three lining conditions over a wider range of hydrogeologic and climatic environments. While it is unlikely th at this would change the overall conclusions regarding the relative risk associated with the three liner conditions (unlined having the highest risk and clay-lined sites having a s ubstantially higher risk than composite-lined sites; RA pgs. 4-39 to 440), it would provide a better indication of the benefits to be gained by installing/mandating different liners.

While the information in RA Tables 4-5 to 4-10 for “unlined”, “clay-lined” and “composite-lined” units is useful for the deve lopment of management options for future CCW facilities, it only considers the specific liner conditions at the specific sites where those conditions actually exist (e.g. the evalua tion of risks for “clay-lined” is only done for sites that are clay-lined). This is appr opriate for evaluating the risk associated with these facilities, however even better insight regarding the effect of the three liner conditions (“unlined”, “clay-lined” and “compos ite-lined” ) could have been obtained by examining all three liner possibilities for a ll sites used in the Monte Carlo study (irrespective of what is currently used) thereby giving an assessment of the risk associated with the three lining conditions over a wider range of hydrogeologic and climatic environments. While it is unlikely th at this would change the overall conclusions regarding the relative risk associated with the three liner conditions, it would provide a better indication of the benefits to be gain ed by installing/mandating different liners.

EPA Response Because they reflect liner conditions in 1995 (the tim e of the Electric Power Research Institute [EPRI] survey), EPA acknowledges that the “all units combined ” risk results may not be representative of current management practices provided that the 2006 DOE/ EPA data is an accurate representation of the nationwide distribution of liner types. To minimize potential misinterpretation of the results, EPA has removed this category from the revised risk assessm ent. EPA now reports groundwater and surface water results separately for each of the 15 management scenarios modeled (see Table 2-3 of the risk assessment). The 15 management scenarios can be seen in Table 4 below. Because the results for each of these management practices accurately reflect the risk s from units managed in that manner, the changing prevalence of liners, units, or wastes will not affect the practice-specific risk results.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

25 

### Table 4. Management Scenarios Modeled

Landfills Surface Impoundments

Liner Type CCW CCW and Coal Refuse FBC CCW CCW and Coal Refuse FBC

Unlined T T T T T N/A Clay-Lined T T T T T N/A Composite-Lined T T T T T N/A
T = Scenario was modeled in the risk assessment N/A = Not applicable (EPA is unaware of FBC wastes handled in surface impoundments)



### 4.3 Alternative Waste Disposal/Codisposal Scenarios

Comment Summary Dr. Rowe expressed concern about whether th e risk assessment considered alternative CCW disposal scenarios. Specifically, codisposal with wastes generating more acidic conditions may significantly impact the mob ility of metals. Dr. Hopkins suggested that the risk assessment should consider CCW benefici al reuse options such as structural fill.

Dr. Kerry Rowe, Queens University However as previously indicated in respon se to Question #2, it is not clear from the documentation how one can validly distinguish between the leachate from conventional CCW and CCWs co-disposed with coal refuse if data is based on the TCLP. One would anticipate that there would be concentrations of metals in the more acidic leachate associated with co-disposal th an in the leachate from a conventional CCW where the pH is expected to be higher.

Dr. William Hopkins, Virginia Polytechnic Institute and State University Once revised, the risk assessment will be an important document for determining what management options for coal combustion wastes are most protective of human and ecological health. With this in mind, it seems odd that one of the most attractive management options, recycling a nd reuse, is not considered in the risk assessment. While it would clearly be far beyond the scope of th is risk assessment to evaluate “beneficial uses” of coal combustion wastes on a broad scale (e.g., wallboard, cement, etc.), this risk assessment not only serves as an important tool for guiding state policies and procedures for disposing of the wastes, but also for using these materials for structural fills.

An increasing number of states consider the use of coal combustion wastes as structural fill material to be a “beneficial use”. Th e National Academy of Sciences encouraged using coal combustion wastes as fill material, but not at the expense of environmental health or in the name of “beneficial use”. To be clear, I believe it is highly desirable to use these wastes for structural fill and as fill material in mines in some circumstances. However, it is clear from documented damage cases (which include at least one superfund site), this draft risk assessment document, and the National Academy of Sciences 2006 report that placement of these materials in landfills can produce undesirable risks if not managed properly. These risks can be mitigated by usingDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

26 appropriate safeguards such as liners. Similarl y, use of these wastes as a structural fill, particularly in large monofills without safe guards, could have adverse effects on the environment. Because structural fills are often categorized as a “beneficial use”, they are often exempt from standards and regulations ( no liners or monitoring requirements, etc.) that provide safeguards from undesirable outcomes. In light of the findings of this risk assessment, the EPA and state regulatory agencies should consider under what circumstances structural filling with coal combustion wastes is safe, and under what circumstances it poses unacceptable risks that outweigh any perceived benefits.

EPA Response The 2003 CCW risk assessment addressed codisposal of coal combustion ash with coal mill rejects (which tend to generate acidic conditions) through the use of the CCW constituent database, which included leachate and porewater samples from codispo sal sites as well as conventional fly ash, flue gas desulfurization sludge, and fluidized bed combustion waste monofills. The risk assessment report states in various places that codisposal was considered, includi ng in Section 1.3.5, where it specifies that one of the three waste types considered is “codisposed CCWs and coal refuse, which are more acidic than conventional CCWs due to sulfide minerals in the mill rejects”.

The studies at Vanderbilt University (U.S. EPA, 20 06c, 2008c) focus on conditions prevalent in CCW sites and were designed to represent current waste disposal scenarios. We extended the risk assessment report with a comparison between risk assessment s ource concentrations and measured concentrations from the most recent report on the Vanderbilt study (see Section 4.1 of this comment response document).

Evaluation of risks from beneficial use of CCW w as outside of the scope of the risk assessment.


### 4.4 Liner Leakage

Comment Summary Dr. Rowe submitted a comment addressing the liner leakage assumptions used in the CCW risk assessment. The composite liner leakage distribution used in the risk assessment was based on empirical data from industrial liners compiled by Tetra Tech (2001). The leakage distribution excluded Tetr a Tech data from landfills with composite liners containing compacted clay layers. The original justification for excluding these types of liners was that water is released from these liners through consolidation and may not represent actual leakage. However, Dr. Rowe argues that this was an incorrect assumption, because compacted clay layers do not release significant volumes of water due to subsequent compaction.

Dr. Kerry Rowe, Queens University The leakage through composite-lined CCW s ites was based on data for a limited number of existing industrial landfill cells (RA Pg . 3-11). Leakage can vary substantially depending on the level of c onstruction quality control and construction quality assurance (CQC/CQA), the wrinkling of the geomembrane when covered, and whether the clay liner component is a compacted clay liner or geosynthetic clay liner (GCL). Finally the service life of a geomembrane is finite (albeit that it may be many hundreds or even a thousand or more years) yet it appears fr om the documentation that the geomembrane was likely to have been intact for the entire period modeled (up to 10,000 years). It is expected that if these factors had been considered the risks associated with the composite-lined sites would have been higher than reported in the RA (Tables 4-5 to 4-DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

27 10). However this would not change the c onclusion that composite-lined sites are superior to clay-lined sites and that these are superior to unlined sites in terms of reducing risk.

In the past few years there has been c onsiderable research suggesting that (a) geomembranes have a finite service life and (b) leakage through composite liners depends significantly on both the nature of th e clay component of the composite liner (CCL or GCL) and the presence of wrinkles wh en the liners is covered (which may cover up to 25% of the liner area and hence this removes composite action from up to a quarter of the total area). Consideration of these f actors (see response to Questions #1 above) would improve the RA and increase the calcula ted risk associated with composite-lined sites compared to that predicted by the RA. However as already noted above, this would not change the conclusion that composite-lined sites are superior to clay-lined sites and that they are substantially superior to unlined sites in terms of reducing risk.

The exclusion of data relating to composite liners with a compacted clay liners (RA, pg.3-11) is NOT considered reasonable. The am ount of water related to consolidation can be shown to be small. Also, composite liner s with a compacted clay liner (CCL) can be expected to leak much more than tho se with GCLs because of the poorer interface transmissivity (many orders of magnitude higher) for CCLs. This approach adopted introduces an unconservative bias to the co rresponding risk assessment for composite lined sites.

Furthermore, the results for composite-lined sites do not appear to adequately recognize a finite service life for the geomembrane component of the liner (as noted earlier). This will result in a consistent underestimate of risk associated with these liners (and hence represents a bias). However this would not ch ange the conclusion that composite-lined sites are superior to clay-lined sites and that they are substantially superior to unlined sites in terms of reducing risk.

The assumption of a clay liner hydraulic conductivity of 1x10-9 m/s (1x10-7 cm/s; RA pg. 3-20) is reasonable provided that the liner had appropriate (a) specifications (in terms of clay content, plasticity, compacted wate r content and density etc.), (b) construction quality control and assurance during constructi on, and (c) it was kept moist (to avoid desiccation) prior to covering with waste. Indeed if these conditions are all met and provided the typically specified hydraulic conductivity of 1x10-9 m/s was achieved at the time of covering (as any well-designed and cons tructed clay liner will) then, following loading by waste in a landfill, consolidation will likely result in a significant reduction in hydraulic conductivity (to say 1x10-10 m/s or possibly as low as 5x10-11 if there is significant applied stress). Thus I consider that the value of 1x10-9 m/s used is actually conservative and likely overestimates the risk associated with well-designed and constructed clay-lined sites.

On the other hand, public comments claim that current practice involves the use of “compacted CCW as liner materials, if any liner is installed” and that “the few CCW landfills that are constructed with clay liners are generally designed with only 2-feet of 10-7 cm/s compacted clay rather than the 3 feet that are assumed ...”(Geo-Hydo, pg.4). If CCW material is used and it does not meet typical clay liner specification, then this should not be modeled as a clay-lined but ra ter as an unlined site. If, for a CCW landfill, the clay liner meets specification and has an a constructed hydraulic conductivity of 1x10-9 m/s then the fact that it may only be 2-feet thick (as opposed to the 3-feetDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

28 assumed) will likely be off-set by the decrease in hydraulic conductivity due to consolidation under the weight of the CCW and the predicted risk is still likely to range between reasonable and conservative (i.e. erri ng on the high side for clay-lined sites).

As previously noted above, the leakage calcula ted as described in the RA (pg. 3-20) is likely to underestimate the leakage through co mposite liners and hence underestimate the risk associated with these liners. That said, properly constructed composite liners are still going to substantially reduce the risks asso ciated with CCW disposal compared to unlined and clay-lined sites and this refine ment will not change the general conclusions of the RA.

... as discussed in response to earlier questi ons above, while variability was considered with respect to MWU characteristics, there does not appear to have been adequate consideration of uncertainty associat ed with composite-lined sites.

EPA Response The composite liner leakage rates used for this ri sk assessment correspond to leakage rates developed for the peer-reviewed Industrial Waste Management Evaluation Model (IWEM) (U.S. EPA, 2002b). The exclusion of data from composite liners containing compacted clay layers was based on an assumption that significant volumes of water can be released from clay due to consolidation from mechanical compaction from the weight of overlying waste. Seve ral literature references di scuss such consolidation water associated with clay liners (Bonaparte et al., 2002; Bergstrom et al., 1993; Gross et al., 1990; Workman, 1993; Feeney and Maxson, 1993). The followi ng simple calculation illustrates that this effect is realistic at least in some cases. Assuming the po rosity of a 2-ft thick clay decreases by 0.05 over 5 years, the resulting flow rate would be 167 L/hectare- day, which is similar to many values recorded by Tetra Tech (2001) for compacted clay liner (CCL) –based composite liners. Several researchers have compared the chemical composition of leachate versus leak detection system water and concluded that a significant fraction of the water is due to conso lidation (Bonaparte et al., 2002; Workman, 1993). Through laboratory consolidation studies, Bergstrom et al. (1993) estimated consolidation water flow rates that exceeded observed leak detection system flows. Although clay liners are clearly thought to release consolidation water, the fraction of leakage a ttributable to consolidation has not been quantified accurately to date. Due to the absence of an appro ach for estimating consolidation flow rates versus leakage, EPA excluded clay liner data from the estima ted distribution of composite liner leakage rates.

Although there is clear literature support for the assu med liner leakage distribution, EPA did evaluate potential changes if leakage rates for CCL-based composite liners had been included in the analysis. Figure 1 compares the resulting distribution of groundwater dilution attenuation factors (DAFs). A DAF is the ratio between the source concentration and the receptor concentration. In this figure, values of 1E+30 represent simulations where the contaminant did no t reach the receptor. As expected, the change in leakage rate does have a significant impact. A larger percentage of simulations showed the contaminant reaching the receptor (18% for the alternative ra tes versus 6% for the original risk assessment distribution). In addition, the DAFs are significantl y smaller for the alternative leakage rates.

Although changing the liner leakage assumptions woul d change the risk assessment results for compositelined facilities, it would not alter the general conclusi on about the relative risk of different liner scenarios (composite lined landfills pose significantly less potentia l risk than unlined landfills). This latter point was emphasized by Dr. Rowe in his review of th e CCW liner leakage and liner scenario discussions.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

29 0 0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8 0.9

1

1.E+00 1.E+03 1.E+06 1.E+09 1.E+12 1.E+ 15 1.E+18 1.E+21 1.E+24 1.E+27 1.E+30

DAFPercentile CCW RA leakage rates Alternative leakage rates

Figure 1. DAF Comparison for different leakage rates.

With respect to the clay liners themselves, the 200 9 risk assessment used the assumption that compact clay liners were designed to have a hydraulic conductivity of 1×10 -7 cm/sec. This is consistent with EPA’s Industrial D Guidance, which states that “clay liners should be at least 2 feet thick and have a maximum hydraulic conductivity of 1×10 -7 cm/sec.” 1 However, clay liners designed to meet a 1×10 -7 cm/sec hydraulic conductivity could perform differe ntly in practice. In Moo-Young et al. (2004), 2 a small set of clay-lined landfills were found to have field hydraulic conductivities ranging from 2×10 -9 to 4.4×10 -8 cm/sec and a small set of surface impoundmen ts were found to have field hydraulic conductivities ranging from 3×10 -6 to 3.2×10 -5 cm/sec. EPA discusses this uncertainty in Section 4.4.3.2 of the revised risk assessment.



### 4.5 Treatment of Nondetect Values

Comment Summary Dr. Basta and Dr. Rowe provided comments on the treatment of nondetect values in the CCW risk assessment. Dr. Basta’s comments fo cused on the assessment’s use of half of the detection limit when a constituent’s concentr ation was not detected. He felt that lower values should have been used, especially for ar senic and mercury, which were frequently not detected but drive exposure pathways. Dr. Rowe suggested that the analysis could use the mercury concentration distribution (as shown in Figure 4-6 in the 2007 CCW risk

1 U.S. EPA (Environmental Protection Agency). 2006. Gu ide for Industrial Waste Management. Office of Solid

Waste, Washington, DC. http://www.epa.gov/waste/nonhaz/industrial/guide/index.htm.
2 Moo-Young , H., B. Johnson, A. Johnson, D. Carson, C. Lew, S. Liu, and K. Hancock. 2004. Characterization of

Infiltration Rates from Landfills: Suppo rting Groundwater Modeling Efforts. Environmental Monitoring and Assessment 96:283–311.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

30 assessment document) to asses s the risks while acknowledging the shortcomings (the likely overly conservative natu re) of such an analysis.

Dr. Nicholas Basta, Ohio State University The “Calculation of average constituent con centrations by site for landfill leachate, surface impoundment porewater, and total ash concentrations” (p. A-5) in the Draft
states:

“During site averaging, any separate waste di sposal scenarios occurring at a site (e.g., non-FBC and FBC ash) were treated as separate “sites” and were averaged independently. This approach is consistent with that used in the 1998 CCW risk analysis. A large number of CCW samples report “below detection limit” values. As in 1998, nondetects were averaged at one-half the reported detection limit.”

However, a large number of CCW samples re ported “below detection limits” for many samples and elements. Two critical elements, As and Hg, had an excessive number of “nondetects” reported. Therefore, there is great uncertainty in the input data for the risk assessment of these elements. This is criti cal because As drives the Human Health drinking water exposure pathway standard, a nd it is very possible that Hg would have driven the Ecological Risk exposure pathway. The analytical laboratory instrumental detection limits are not reported. Cursory insp ection of the data in Appendix A-2 suggests improved analytical methods should have been u sed to achieve lower detection levels that would have decreased or eliminated the larg e number of nondetects. Analytical methods that achieve lower detection limits can have an impact on the risk calculation for As. It is very possible that leachate concentration could be much lower (tenfold or more) than half the high detection levels reported. Modern analytical methods can achieve detection limits of < 0.001 mg/L for As. These methods should be used to analyze CCW leachate or extract data for CCW landfill risk assessment.

Dr. Kerry Rowe, Queens University It is acknowledged that a large number of nondetects, as observed for mercury, poses a considerable challenge for a risk assessment. The use of half the detection limit as the concentration for “nd”s (RA, pg. 4-34) to represent the concentration may give unrealistically high impacts (as is suggested by the data in Figure 4-6). However to ignore mercury altogether, especially given th at concentration may increase in the future (RA, pg. 4-34) is also questionable. At the ti me the study was conducted, an alternative to ignoring mercury would have been to do anal yses with their concentration distribution (as shown on Fig. 4-6) and assess the risks while acknowledging the shortcomings (likely over-conservative nature) of the analysis. If th e risk was low (below criteria) then one could easily conclude that mercury was not a problem. If the risks were significantly above criteria then this would highlight the need for more study of this issue.

EPA Response As stated in Section 4.3, the leachate concentration was one of three input parameters that the model was most sensitive to. The upper percentiles of leachate values, not the nondetects, are responsible for the upper percentiles of risk. More than half of the sam ples in the CCW constituent database had detectable levels of arsenic. Thus, for arsenic and other metals with a significant number of detectable measurements, it is the detected values that genera te the upper-percentile risk levels. Therefore, from a statistical standpoint, the replacement of these nondet ects with half the detection limit or some other suggested value will not change the 90th percentile risk levels generated by the risk assessment.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

31 Although the treatment of nondetects could change th e 50th percentile values reported, there is no better option available to EPA without great time and expense. Consider the two most extreme options. The first would be to set all nondetects right at the detection lim it. Because this is one of the most sensitive inputs, such a change would likely significantly raise the 50t h percentile risk values, but would more assuredly be protective of human health and the environment. However, it is likely that at least some nondetect samples are actually much lower, resulting in an over estimate of risk. Alternatively, setting all nondetects to zero would be even more inappropriate. Given th e high number of measurements for arsenic and other metals (excluding mercury), it is unlikely that the constituents are completely absent from all the remaining samples. In fact, there were many cases wh ere samples taken from the same facility had both detects and nondetects, making it almost certain that the constituent was present at some level in the nondetect samples. Given the unlikeliness of either ex treme scenario being accurate, EPA determined that there would be some distribution of concentrations between zero and the detection limit. Because the midpoint of the two extremes of such a distribution would be half the detection limit, EPA determined that replacing nondetects with half the detection limit would be appropriate. This is consistent with both the Risk Assessment Guidan ce for Superfund (RAGS) Part A (U.S. EPA, 1989) and with EPA Region 3 Guidance on Handling Chemical Concentration Data Near the Detection Limit in Risk Assessments (U.S. EPA, 1991).

The case for mercury (and thallium and antimony fo r surface impoundments) is very different. Here, virtually all of the measurements were nondetects. Fo r this reason, mercury was eliminated from the analysis to eliminate the potential (and unknown) bi as. Thallium and antimony we re likewise eliminated from analysis in the surface impoundment scenarios. Cl early, the risk assessment with respect to these constituents would be enhanced by new waste measur ements generated by more sensitive instrumentation and carefully designed studies (like the Vanderbilt/EPA studies). Such data were generally not available during the initial CCW data collection effort. The i ssue of mercury, thallium, and antimony is discussed in Section 4.4.3.1 of the revised risk assessment.



### 4.6 Effect of W aste Compaction

Comment Summary Dr. Hopkins suggests considering how waste compaction would reduce water infiltration. He feels that this would be a useful comparison with the current assessment.

Dr. William Hopkins, Virginia Polytechnic Institute and State University 5) Consideration of how compaction of wastes in the landfill would reduce water infiltration. The report does a good job of demonstrating the importance of liners in preventing undesirable migration of constitu ents from the site. However, compaction technologies reduce water penetration through the materials, rendering them less likely to produce leachate. If the data are available, th is would be a useful comparison (similar to the lined vs. unlined comparison) to demonstrate how compaction reduces risks by reducing leachate production.

EPA Response EPA acknowledges this comment and has addressed it by describing it as a modeling uncertainty in Section 4.4.2 of the revised risk assessment:

“The source model did not consider potentia l compaction of CCW waste over time. Such compaction could decrease the hydraulic conductivity and the associated water infiltration. However, no readily available data were identifie d to support an analysis of the influence of CCW compaction on infiltration rates. The cu rrent approach would tend to overestimateDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

32 infiltration rates compared to a model that w ould adjust the hydraulic conductivity over time due to compaction. EPA believes this is an appropr iately conservative assumption given the lack of the information needed to accurately m odel the effects of waste compaction.”


### 4.7 HELP Infiltration Rates/Cover Assumptions

Comment Summary Dr. Rowe questioned the use of the HELP model in light of recent criticisms and expressed concerns that the limitations of this model and the potential implications of its use have not been adequately addressed in the document.

Dr. Kerry Rowe, Queens University
The use of HELP to model infiltration for la ndfills has been the subject of some recent criticism in the literature however the lim itations of this model, and the potential implications of its use, have not been adequately addressed in the document.

Any assessment of infiltration through a la ndfill cover must consider the long-term performance of the cover (after the effects of many seasons of climatic exposure are considered). Thus, for example, the bulk hydraulic conductivity of a compacted clay liner in a cover is likely to increase by orders of magnitude due to desiccation as a result of temperature cycles in most climates (both hot and cold). If a GCL is used as an alternative, its hydraulic conductivity may also increase (in some cases by several orders of magnitude) due to cation exchange with th e adjacent “soils” and desiccation. It is not clear that either of these issues were considered.

EPA Response EPA has adopted the use of the HELP model as fundament al source of unlined and clay-lined infiltration rates for landfills for over 14 years and recognizes that there are limitations in using HELP. However, the HELP model has been tested, verified, and used ove r many years. This is well discussed in the HELP model documentation, as well as the EPACMTP Para meter/Data Background Document (U.S. EPA, 2003a) as discussed and referenced in Secti on 3.4.2.1 of the revised risk assessment.



### 4.8 Scenarios Where the Water Tabl e Is Above the Disposal Unit

Comment Summary Dr. Rowe expressed concern about the disconn ect between the fact that the assessment modeled only disposal sites above the water table and the statement in the report that many CCW disposal sites are constructed within the saturated zone. The commenter did acknowledge that modeling disposal sites be low the water table would only strengthen the assessment’s conclusion rather than change it.

Dr. Kerry Rowe, Queens University Concern was raised in the public comments a bout the modeling only considering disposal sites where the bottom of the facility is a bove the water table while claiming that CCW landfills are “often constructed in mines a nd quarries ... within the saturated zone” which “will have the effect of expediting leaching of contaminants from the waste to levels consistent with levels found in surface impoundments” (Geo-Hydro, pg. 4). If this is indeed the case and if active measures are not en forced to control the head in the landfill to below that outside the landfill (hydraulic c ontainment), then the risk of these sites will indeed be higher than modeled. However given that the general conclusions from theDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

33 study are that the risks for unlined and clay- lined sites already exceed the risk criteria, this would only serve to emphasize that point and would not change the general conclusions.

EPA Response EPA acknowledges that EPACMTP does not currently handl e the scenario where the water table is above the disposal unit, and has added a statement to m odel uncertainty section of the risk assessment report (Section 4.4.2) clarifying this limitation of the analysis:

“The landfill source model and EPACMTP assume that the source is above the water table. However, some actual CCW dispo sal units do extend below the water table. Because waste intersecting the saturated z one may increase groundwater concentrations, the approach may underestimate risk in some cases. However, including this effect would strengthen a general conclusion of the analysis that risks above the risk criteria exist in some cases with unlined and clay lined CCW landfills.”

As recognized by the commenter, the general conclusi ons regarding the risks of unlined and clay-lined landfills would not change if disposal below the water table were modeled.


### 4.9 Impact of Increased Mercury Captur e with New Future Air Emission Control Regulations

Comment Summary Dr. Hopkins, Dr. Rowe, and Dr. Vorhees recommended that EPA address the future increase in mercury levels in CCW that w ill result from mercury capture from flue gas under new emission control regulations. Dr. Rowe and Dr. Vorhees pointed out that the recent Vanderbilt study should pr ovide data that could be used to expand the risk assessment in this area.

Dr. William Hopkins, Virginia Polytechnic Institute and State University Because Hg is volatilized during combustion, it is not surprising that Hg was not detected in many circumstances in the solid wastes. However, this question raises an important point: Will the conclusions of the risk assessment remain similar once Hg is more effectively captured and introduced into th e solid wastes? This will not only be a tremendous challenge for the power industry and for the corporations that recycle these materials, but also for the USEPA as they rec onsider the risks from a very different waste stream. Importantly, the amount of Hg that w ill be captured in the future and introduced into the solid waste stream should be predictable. Based on these predictions, I recommend that these risk assessment models be run based on current as well as projected waste composition (by manipulating Hg concentrations in the models). This will be invaluable for addressing management of coal combustion wastes when Hg concentrations increase in the near future.

Dr. Kerry Rowe, Queens University If more work were to be done today there is at least the data from the Vanderbilt study however it would appear that if more work is to be done it should also look at the implications of flue gas mercury controls.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

34 Dr. Donna Vorhees, The Science Collaborative EPA’s approach to treating the high frequency of concentrations reported below detection limits is reasonable, particularly given its ac knowledgement that more work is needed to resolve questions about mercury because of its tendency to bioaccumulate in fish and because mercury concentrations in CCW might increase as flue gas mercury controls are installed on coal-fired power plants. EPA indicates that mercury concentration and leachability from CCW can vary as a function of flue gas controls and that “additional work is underway in this area” (page 4-41). What is not made clear is when this additional work will be comp leted and how it will be used to reassess risk from WMUs used to dispose of CCW?

EPA Response The scope of the analysis was to evaluate current waste disposal conditions, no t potential future changes associated with emissions controls. EPA added a statem ent to Section 4.4.3.1 of the risk assessment report clarifying that analysis of mercury emission controls was outside the scope of the current risk assessment:

“Mercury is of particular interest because it is the only constituent with significant concern through the fish consumption pathway, a nd because there is the potential for mercury concentrations in CCW to increase as flue gas mercury controls are installed on coal-fired power plants in response to the Clean Air Interstate Rule (CAIR) and the Clean Air Mercury Rule (CAMR). However, analysis of the effect of me rcury emission controls was outside the scope of the risk assessment, which was to evaluate current waste disposal conditions, not potential future changes due to emission controls.”

The same section also describes the recent EPA studies conducted by Vanderbilt University:

“Finally, U.S. EPA (2006c) and preliminary results of ongoing EPA studies (e.g., U.S. EPA, 2008c) suggest that both mercury levels and merc ury leachability in CCW can vary depending on the flue gas mercury controls used at a power plant.”



### 4.10 Use of Porewater Data for Surf ace Impoundment Source Concentrations

Comment Summary Dr. Rowe expressed concern about estimating surface impoundment source concentrations using measured concentrati ons in pore water collected from underlying sediments. The pore water concentrations are lik ely lower than concentrations in the bulk surface impoundment water, because they reflect any attenuation between the surface impoundment and the measurement point.

Dr. Kerry Rowe, Queens University While I would have preferred leachate data for the surface impoundments based on analysis of the actual leachate itself, the data obtained from pore water sampling of soil from below the impoundments (RA, pg. 4-32) is considered as a reasonable alternative as long as the limitations are recognized. It should be recognized that the pore water squeeze chemistry may have been impacted by geoche mical reactions (e.g. precipitation, sorption) during migration through the “soil” above the sample from which the pore water was extracted. Since these interactions will vary from one “soil” to another (depending on the mineralogy of the soil), the pore squeeze data will reflect this variab ility but will have a bias to lower values. Thus while the mode ling does consider retardation mechanism as contaminants migrate through the soil, the use of a source term that is artificially lowDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

35 because of sorption/precipitation that occurre d in the soil above which the sample was taken would lead to a potential underestimate of impact.

Dr. William Hopkins, Virginia Polytechnic Institute and State University The quantity of data is not sufficient. Surface water concentrations need to be considered 1) within the impoundments and 2) downstr eam from impoundments (i.e., direct surface water contamination, independe nt of subsurface flow).

For surface impoundments, pore water data seem appropriate for estimating subsurface risks, but not for the two most important exposure pathways (direct exposure in surface impoundments and downstream from surface impoundments). To complete these analyses, a minimum of two other types of information are required. First, actual concentrations of trace elements in coal co mbustion wastes are needed to assess risk to animals (deer, dabbling ducks, grazing amphibian larvae, certain invertebrates) that directly ingest the solid waste materials wh en foraging in or around the margins of impoundments. These data are readily available, making this portion of the risk assessment highly feasible. Second, surface water concentrations in the impoundments themselves should be used to assess risks to animals that drink water from impoundments (a wide array of invertebrates and vertebra tes [especially birds and mammals]) or deposit their eggs directly in the water (amphibi ans, invertebrates). These data are likely available, at least from discharge points which should represent the low end of actual concentrations in the impoundments. To assess risks in downstream sites affected by discharge, different dilution factors can be a pplied to represent a range of site-specific scenarios. In addition to calculating resultant concentrations afte r dilution, continual loading and deposition need to be considered in downstream habitats based on the duration of site operation and discharge.

EPA Response The porewater concentrations were used for estimati ng fluxes to groundwater. The use of porewater data for this purpose is most appropriate for the CCW risk assessment because porewater more closely represents the leachate seeping from the bottom of the impoundment than do impoundment water samples. EPA has provided a clarifying statement in Section 4.4.3.1 of the revised risk assessment:

“Surface impoundment leachate is represented by pore water measurements taken beneath actual impoundments, which should more closely represen t the leachate seeping from the bottom of the impoundment than would bulk surface impoundment w aste concentrations. The pore water is in direct contact with the waste, so these concentr ations should typically be at least as great as concentrations in the bulk surface impoundment.”

Also Section 3.5.1 has been revised for clarity as follows:

“Accordingly, the surface impoundment source-te rm model also assumes that the leachate concentration is constant during the impoundment operational life and equal to the concentration in the porewater in the sediments at the bottom of the impoundment.”


### 4.11 Modeling Surface Impoundments Only Before Closure

Comment Summary Dr. Hopkins pointed out that many surface impoundment wastes are left in place at the time of closure, so that the waste behave s more as a landfill than a surface impoundment.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

36 The risk assessment considered releases from surface impoundments only during active operation.

Dr. William Hopkins, Virginia Polytechnic Institute and State University Consideration of how surface impoundments are actually managed. In many cases the coal combustion wastes are left in place at the time of closure (analogous to filling a landfill). Thus, risk does not end when the slurry pumping ceases, which is in direct contrast to the assumptions of the current risk assessment model. Instead, the type of risk transitions from that modeled in the surface impoundment to that which is more analogous to a capped landfill. Again, this inaccurate assumption underestimates risk.

EPA Response The commenter is correct in regard to the clean closure assumption for the surface impoundment model. EPA assumes that risks via groundwater are much higher during surface impoundment operation because the higher hydraulic head drives leachat e into the underlying soil with gr eater force than infiltration after the impoundment is covered and closed. This higher head is assumed to result in a greater flux of contaminants to groundwater during the active life of the surface impoundment, especially in unlined units.

This has been clarified in the revi sed risk assessment, including a discu ssion of this issue under the model uncertainty section (Section 4.4.2): “As described in Section 3.5.1, the surface impoundment model treats a surface impoundment as a temporary waste management unit with a set operational life. At the end of this life, clean closure is assumed; all wastes are removed and ther e is no further release of waste constituents to groundwater. Although this simplifying assumption is not consistent with the practice to close CCW surface impoundments with wastes in place, and it limits the length of potential exposure, the peak annual leachate concentrations on which the CCW risk results were based are not likely to be affected. Releases to groundwater ar e much higher during surface impoundment operation because the higher hydraulic head in an ope rating impoundment drives wastewater into the underlying soil with greater force than inf iltration through the impoundment cover after the impoundment is closed. This higher head results in a greater flux of contaminants to groundwater during the active life of the surface impoundment, esp ecially in unlined units. Thus, even if the post-closure period were modeled, the correspondi ng results would not be as high as the peak annual leachate concentrations used in the analysis.”

In addition, EPA notes that the arrival times fo r many of the unlined and clay-lined surface impoundment risks are either within the 75-year operating life assumed in the model, or just beyond it. As seen in Table 4-7, the median arriva l time for peak boron and selenium VI risk is 74 years. Thus, even with further leaching occurri ng during year 75, the risks in the typical model run do not increase. Thus, EPA believes it is unlikely that waste-in-place management would significantly affect the model results.


### 4.12 Modeling Landfills Only After Closure

Comment Summary Dr. Hopkins suggested that risk from operati ng landfills should have been considered, because the greatest risks may occur during s ite operations when the waste products are directly exposed to wind and rain.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

37 Dr. William Hopkins, Virginia Polytechnic Institute and State University It is incorrect to assume that risk begins when the landfill closes. In reality, the greatest risks likely occur during site operations when the waste products are directly exposed to wind and rain. This is a simple, yet serious flaw in the risk assessment that obviously underestimates overall risk.

EPA Response EPA acknowledges this comment. The simple landfill m odel used in the CCW risk assessment did not model leaching from landfills before closure, as the w aste is placed. This issue and other peer-review comments related to exposure scenarios during landfill operations are addressed in Section 5.4 of this document.



### 5.0 COMMENTS RELATED TO EXPOSURE SCENARIOS The peer review comments relating to exposure scen arios are grouped into five subcategories: 1. Direct surface impoundment exposure 2. Indirect exposure through game species (e.g., deer, waterfowl) 3. Exposure to surface impoundment effluent 4. Exposure during landfill operation (before closure) 5. Elimination of pathways during screening analysis.


### 5.1 Direct Surface Impoundment Exposure

Comment Summary Dr. Hopkins pointed out that the risk assessment does not consider direct exposure to active surface impoundments. Such exposure can occur, for example, when wildlife are attracted to impoundments for uses such as dri nking, feeding, and swimming, can lead to exposure through transdermal or direct ingestion pathways. There is also a potential for trophic exposure increases as other organisms eat animals or plants exposed to surface impoundment waste.

Dr. William Hopkins, Virginia Polytechnic Institute and State University The National Academy of Sciences concluded that coal combustion wastes pose the greatest risk to humans and the environment when they contact water in the disposal environment. Thus, by definition, surface impoundments should pose greater risks than most landfills because the use of surface impoundm ents requires that the wastes be mixed with water before being pumped into the im poundments. Overall, the findings of the risk assessment are consistent with this. That is, risks are higher for surface impoundments than for landfills. However, th e risk assessment grossly unde restimates risks from surface impoundments to humans and the environment b ecause it fails to consider the two most important exposure pathways for invertebra tes, fish, wildlife, and humans. For surface impoundments, the risk assessment only rec ognizes leaching as a re lease mechanism and 1) groundwater-to-drinking water and 2) groundwater-to-surface water as exposure pathways. This is incorrect. While leaching in to groundwater is clearly a concern at these sites, the predominant exposure pathways rela te to 3) direct exposure to the slurried wastes in the impoundments and to 4) di rect exposure to the effluent from the impoundments discharged offsite. I highlight th e problems with overlooking these routes of exposure below.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

38 Direct Exposure Onsite. Plants and animal s can experience direct exposure to coal combustion wastes when they inhabit or periodically utilize surface impoundments. Surface impoundments often attract wildlife from adjacent areas for a variety of purposes, including foraging, drinking wate r, and reproduction. Examples include game (e.g., deer, migratory waterfowl, and racc oons) and non-game species (migratory passerine birds, wading birds, amphibians, a nd reptiles). For these animals, exposure to trace elements in coal combustion wastes can occur by at least four mechanisms:

A) Drinking water from the impoundments,

B) Direct ingestion of the wastes,

C) Trophic exposure caused by eating other animal inhabitants or ingesting plants growing in or around the impoundment, and/or

D) Transdermal exposure.

Drinking water is a widespread mechanism of exposure and is a significant risk to most taxonomic groups since the impoundments can cont ain large quantities of fine, suspended particulate matter (e.g., fly ash) and high c oncentrations of dissolved trace elements resulting from the slurry. Several important gr oups of organisms ingest sediments from the bottom of impoundments, perhaps most notably, dabbling ducks (e.g., mallards, teal, pintails, etc.) and larval amphibians (i.e., fr ogs and toads). Aquatic plants and algae often comprise a significant portion of the diet for a variety of these animals, as well as aquatic turtles which frequently become residents of surface impoundments. Transdermal exposure is possible for all species that co me in contact with the water and coal combustion wastes in and around surface im poundments, but is especially problematic for species with highly permeab le skin (e.g., amphibians).

From an ecological perspective, all of th e groups of organisms mentioned above are important. It is possible that such sites ac t as population sinks in the landscape for species such as amphibians and certain birds (e.g., gr een herons and red-winged blackbirds) that use them in large numbers for reproductive purposes. In other cases, animals that drink or forage at the sites can transport accumulated pollutants to adjacent habitats where their body burden becomes a bioavailable exposure risk to their predators.

EPA Response The direct exposure of ecological receptors to CCW surface impoundment wastewater was addressed in the screening risk assessment, where direct expos ures to both surface impoundment wastewater (using data from the 1998 risk assessment) and surface impoundm ent porewater were considered and compared. Although the porewater risk estimates were intende d to represent surface waters contaminated with groundwater, they were based on u ndiluted porewater concentrations, and thus could represent direct exposure to wildlife as well. They also include tro phic exposure increases (i.e., bioaccumulation up the food chain). Results from the 1998 CCW risk assessmen t, which were described in the 2002 CCW risk screening document, support the conclusion that eco logical risks from direct exposure to surface impoundment wastewater are significant, which agrees with the commenter. These conclusions can be further supported with damage inci dents in published literature. As a result, EPA did not evaluate these risks further in the full-scale assessment, consideri ng the screening results and damage cases to be sufficient evidence that th ese risks are significant. EPA has integr ated the 1998 direct exposure results along with the screening risk results into the revised risk assessment document (direct exposure results are in Section 4.2.1, and screening risk results are in Section 3.2.4). EPA has also summarized damage casesDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

39 (see Sections 4.2.5 and 4.4.1 of the revised risk assessment document) to more explicitly document and support these conclusions in the risk assessment.


### 5.2 Indirect Exposure Through Game Species (e.g., Deer, Waterfowl)

Comment Summary Dr. Hopkins expressed concern that the risk assessment does not consider potential indirect exposure to humans through game sp ecies that may have been exposed to surface impoundment waste (e.g., deer drinking surface impoundment water).

Dr. William Hopkins, Virginia Polytechnic Institute and State University From a human health perspective, game speci es are particularly concerning. For example, a study on landfilled coal combustion wastes w ith improper cover demonstrated that deer are attracted to the wastes (and ingest it), presumably because of its salt content (Sample and Suter, 2002). Thus, in a surface impoundment setting, deer not only drink from the impoundments, but may also ingest the materials directly from the impoundment margins. While it may be possible to regulate hunting in areas near surface impoundments to protect humans from such e xposure, it is less likely that hunters (and their families) could be protected from the h azards of eating waterfowl that have used surface impoundments. Migratory waterfowl often use surface impoundments and in some cases (depending on where the impoun dment is located along their migratory range) could remain on site for considerable periods of time. When these waterfowl leave the impoundment, they become a route of exposure for hunters in other (perhaps distant) localities.

EPA should consider this important mechanis m of exposure in their risk evaluation so that appropriate guidance can be provided to the power industry. If a landfill was left uncapped and a wide variety of wildlife species used the site (e.g., eating plant material growing in the wastes, burrowing in the wast es, nesting in the wastes), it would not be acceptable. Minimum cover standards are in place to prevent such ecological problems. Paradoxically, surface impoundments can actually attract wildlife and promote exposure to high concentrations of trace elements in bioavailable forms. These exposure scenarios are not considered in the current draft ri sk assessment and as a result, the assessment grossly underestimates exposure and resultant risks.

EPA Response We agree that this human exposure scenario was not a ddressed in either the screening analysis or the fullscale CCW risk assessment, and that risks may be sign ificant if humans consume wildlife contaminated by CCW constituents. The scenario uncertainty section of the risk assessment (S ection 4.4.1) has been revised to add a clear statement that this pathway was not considered:

“In addition, not all possible exposure pathways we re evaluated. For example, the risk assessment did not consider potential indirect exposure to hum ans through game species that may have been exposed to surface impoundment waste (e.g., deer drinking surface impoundment water). This represents a potential uncerta inty in the analysis.”DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

40 

### 5.3 Exposure to Surface Impoundment Effluent

Comment Summary Dr. Hopkins emphasized that the risk assessment does not consider potential exposure to surface impoundment effluent even though many CCW waste management systems continuously release wastewater into downstre am habitats. Such exposure could be direct for animals living downstream (e.g., fish) or indirect (e.g., humans catching fish downstream).

Dr. William Hopkins, Virginia Polytechnic Institute and State University Direct Exposure Offsite. Plants and animal s can experience direct exposure to coal combustion wastes downstream from surf ace impoundments. Depending on how the site is engineered, surface water in impoundment s can be recirculated or discharged downstream. Discharge of surface waters into reservoirs can result in significant adverse ecological impacts such as the classic case study at Belews Lake, NC where most fish species (19 of 20) were extirpated. Today, contaminated surface waters are commonly discharged into streams and rivers where dilutio n factors are large and offsite transport is typical. While concentrations of trace elemen ts in surface water are required to meet discharge permit standards, continual loading of trace elements to downstream habitats is a significant concern for human health and the environment. In addition, if not managed properly, rain events can result in resuspensi on of coal combustion wastes and the direct discharge of suspended particulates into downstream habitats. These problems can be further exacerbated in large rivers that r eceive discharge from surface impoundments at multiple power plants, creating cumulative e xposure risks in downstream habitats for human and ecological receptors.

Although the environmental impacts of offsite discharge into lakes/reservoirs are well known, the impacts to streams and rivers have received less attention. However, recent work near Atlanta, GA demonstrated that surface impoundments represent an important source of bioavailable trace elements including As, Se, and Cd to the Chattahoochee River, an important water source and exposure pathway for humans and wildlife (Loeffler-Peltier, 2008). The biological effects of such discharge to rivers and streams are not well studied, but invertebrates and microbial communities have been adversely impacted by surface impoundment discharge (Reviewed in National Academy of Sciences 2006 Report, pg 89-90). Likewise, discharge of trace elements from impoundments to rivers can result in artific ial selection for more antibiotic-resistantcommunities of micro-organisms (Stepanauskas et al., 2005). This latter finding has farreaching human health implications.

Because many surface impoundments continually release contaminated surface waters into downstream habitats, direct discharge is an important release mechanism that must be considered in EPA’s risk assessment process. For humans, the most likely exposure pathway will be ingestion of fish or other commonly eaten vertebrates (e.g., turtles, ducks, or muskrats) that inhabit downstream s ites. However, recreational use of impacted streams and rivers could also result in si gnificant contact with water and sediments, creating multiple exposure pathways. For fish a nd wildlife, particular attention should be devoted to species that utilize areas that favor deposition of trace elements and fine particulates, including areas of reduced fl ow (deep holes, oxbows, etc.) that are also commonly used for recreation by humans (fishing, swimming, etc.).DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

41 This represents the greatest flaw in the ri sk assessment. Direct exposure in surface impoundments and downstream from surface impoundm ents represents the greatest risk pathway for humans and the environment, yet they are not considered in the assessment. In other words, subsurface flow predicted based on physical and chemical processes only represents a small portion of the total risks posed by these waste management units. Direct surface transport from impoundments and dilution factors need to be considered. See comments above (Major Criticism).

EPA Response Resource Conservation and Recovery Act (RCRA) wast e disposal risk assessments do not address direct discharges from impoundments to surface waters becau se they are regulated as permitted point source discharges under the Clean Water Act by EPA’s Office of Water. Thus, this exposure scenario falls outside of the scope of the risk assessment. However, a discussion of this uncertainty has been added to the scenario uncertainty section of th e risk assessment (Section 4.4.1):

“Because this risk assessment addressed CCW disposal under RCRA, it did not include risks from the direct discharge of wastes into waterbodies, which are regulated under the Clean Water Act. Although not relevant fo r the management of RCRA waste disposal, EPA recognizes that CCW surface impoundment effluent may pose additional risks”


### 5.4 Exposure During Landfill Op eration (Before Closure)

Comment Summary Dr. Hopkins pointed out that the risk assessme nt considered landfill conditions only after closure. This approach excludes potential pathways such as exposure to airborne contaminants associated with dry handled waste and direct ingestion of waste by wildlife.

Dr. William Hopkins, Virginia Polytechnic Institute and State University ... the risk assessment should include the following:

1) Consideration of the two most important release mechanisms an d resulting exposure pathways associated with surface impoundments (see comments [Major Criticism] above regarding exposure to contaminants in su rface impoundments and in downstream sites).

2) Consideration of the inhalation pathway for human health risks in a landfill setting. This is a significant problem when the wastes are being predominately dry handled, and can be remedied by adding modest amounts of moisture to the wastes and by limiting the amount of the materials that are exposed to air over the protracted life of an open landfill. Failure to control airborne transport of th e wastes during site operations will result in increased risks for workers and residents in the areas adjacent to transport routes and landfills. Although this concer n was raised repeatedly to the National Academy of Sciences by concerned citizens who could provide concrete evidence of airborne transport (e.g., thin layers of ash covering their cars, homes, etc.), the committee did not evaluate this because it was beyond the scope of their study. However, it is within the scope of this comprehensive risk assessment and should be included.

The exposure distributions do not include the wide array of organisms that are exposed onsite while the waste management units (especially surface impoundments) are active. Because these sites can be active for decad es and this period may represent the greatestDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

42 exposure risk to wildlife (especially in surface impoundments), the current distributions grossly underestimate actual exposures.

Consideration of the starting point for the source term in the landfill model. It is incorrect to assume that risk begins when the landfill closes. In reality, the greatest risks likely occur during site operations when the waste products are directly exposed to wind and rain. This is a simple, yet serious flaw in the risk assessment that obviously underestimates overall risk.

EPA Response As previously discussed in Section 4.12 of this comment response document, the simple landfill model used in the CCW risk assessment do es not handle the complexities of modeling leaching or fugitive dust emissions during landfill operation. EPA acknowledges that the current risk assessment does not evaluate the inhalation pathway. EPA's 1998 risk assessment evaluated this pathway by modeling chronic exposures to constituents in airborne CCW. Howeve r, that assessment did not evaluate acute exposures from particulates. To correct this deficiency, EPA conducted a screening assessment of fugitive dust (U.S. EPA, 2009b) and will address such exposure thr ough dust controls specified in the regulations. With respect to direct ingestion and contact of dry handled wastes by wildlife and plants, the CCW constituent screening analysis (U.S. EPA, 2002) did recognize this exposure pathway:

“Similarly, in landfills, there is a chance that r eceptors could incidentally ingest some whole waste. For example, a white-tailed deer or a vole could ingest whole waste as it browses or forages in the landfill. Small mammals, such as voles, are preyed upon by higher trophic levels, with the potential for bioaccumulation of metals up the food chain.”

In addition, the screening analysis found that boron did pose direct exposure risks to plants:

“ . . . boron exceeds the risk criteria for direct exposure to contaminated soil. This agrees with previous studies noting plant toxicity to boron in coal ash (e.g., Carlson and Adriano, 1993). Because of the magnitude of this screening ex ceedance (3.5 times the risk criteria) and the existence of documented damage cases, we do not believe additional modeling of this pathway is needed to confirm that boron is of concern to pl ants grown in soil contaminated with coal ash.”

The direct contact risks for landfilled CCW risks can be further supported with damage incidents in published literature, such as studies showing impact s of CCW on amphibians at DOE’s Savannah River Plant (Hopkins et al., 2006). Section 4.2.5 of the fi nal risk assessment report discusses published damage cases to document the current knowledge about these ecological exposures.


### 5.5 Elimination of Pathways During Screening Analysis

Comment Summary Dr. Vorhees requested more documentation of th e screening analysis and justification for eliminating certain exposure pathways from the risk assessment.

Dr. Donna Vorhees, The Science Collaborative EPA’s Screening Analysis. EPA conducted a 2002 screening analysis in which it determined the subset of exposure pathwa ys and CCW constituents that would undergo full-scale evaluation in the risk assessment . Figure ES-1 summarizes results of the exposure pathway screening, and Table 2-3 summarizes results of the CCW constituent screening. No discussion is provided about wh y exposure pathways were eliminated. IDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

43 tried to determine the reason, but I could not find the screening analysis document on the website provided in the risk assessment (i.e., http://www.epa.gov/epaoswer/other/fossil/inde x.htm). The risk assessment needs to include at least a summary of these analyses. Also, why is poultry not among the possible animal product pathways? Why is there no dermal contact with drinking water, e.g., during showering? EPA provides guidance fo r assessing dermal exposure to metals in water (See Chapter 3 in EPA 2004a).

EPA Response
The 2002 Constituent Screening for Coal Combustion Wastes (U.S. EPA, 2002a) report was not available to the peer reviewers. However, it has been ma de available in the docket at www.regulations.gov (Document #: EPA-HQ-RCRA-2006-0796-0470). The screen ing report was meant to focus EPA’s efforts towards the pathways and constituents that posed the greatest potential for risk, in particular the constituents for the full-scale risk assessment of the drinking water and surface water pathways.



### 6.0 COMMENTS RELATED TO HU MAN EXPOSURE AND TOXICITY The peer review comments relating to human exposure and toxicity are grouped into seven subcategories:

1. Combined chemical effects, additiv e risk, and multiple pathway exposures 2. Toxicity values 3. Target cancer risk 4. Defining the exposed population 5. Exposure factor assumptions 6. Child/infant exposure assumptions 7. Correlations in input variables.


### 6.1 Combined Chemical Effects, Additive Risk, and Multiple Pathway Exposures

Comment Summary Dr. Rowe questioned whether the possible interaction of people consuming both contaminated fish and drinking water from the same waterbody source was considered and where this was documented.

Dr. Vorhees questioned why a cumulative risk evaluation was not attempted for simultaneous exposure to CCW constituents. Specifically she suggests that potential cumulative exposures of concern could be iden tified by a review of EPA’s existing data related to relative proportion of CCW cons tituents at WMU (waste management unit) locations as well as predictions of CCW cons tituent concentrations in drinking and surface water near WMU locations. Additionally, she noted that cumulative risk needs to be reported from the perspective of the recep tor and that cumulative exposure should be considered for all 21 constituents with risks exceeding chosen benchmark values.

Dr. Hopkins noted that it is not appropriat e to treat each cons tituent separately because they exist in a mixture in the waste stream and some constituents are known to interact. He suggests that this at least be addressed by modeling additivity, although this would underestimate synergistic interactions. Add itionally, he notes that multiple exposureDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

44 routes should be considered with the most likely combinations used for the risk
assessment.
Dr. Kerry Rowe, Queens University The study considered people consuming fi sh from the water body and drinking water from groundwater (without considering interacti on). It is not clear why the potential for people both consuming fish and drinking wa ter from the water body was not considered. This would seem entirely possible especially if people had a residence/cottage adjacent to the water body. Was this possibility examined and if so where is the documentation?

Dr. Donna Vorhees, The Science Collaborative
Cumulative risk refers to the combined risk from exposure to multiple stressors, in this case, CCW constituents. Given that exposu re to CCW constituents might occur simultaneously and some might jointly cause similar toxicity, it is important to evaluate the potential for cumulative risk. EPA has tackle d this topic repeatedly in recent guidance (EPA 2000, 2003, and 2007), so it is puzzling that EPA provides no reason for why it did not attempt a cumulative risk evaluation except to say that such an evaluation is difficult.

In Superfund risk assessments, cumulati ve risk is often addressed by summing HQs across constituents and if the sum is greater than 1, examining the underlying assumption of additivity more closely. However, as noted in at least one set of public commentary, one cannot simply sum 90th percentile HQs across different CCW constituents because this sum might reflect a cumulative exposur e to which no one is exposed in reality. Specifically, the 90th percentile exposure leve l for one CCW constituent might not occur in the same place and time as the 90th percentile level for another CCW constituent. However, EPA does have data that indicate the relative proportion of CCW constituents at individual WMU locations and predicti ons of CCW constituent concentrations in surface water and drinking water near each WMU over time, so it could review these results to check for potential cumulative exposures of concern. In its evaluation of cumulative risk, it is important that EPA report findings from the perspective of the receptor: what combination of CCW constituents will he or she be exposed to within a timeframe where they might collectively cau se harm? This approach is particularly important given the 10,000 year modeling timeframe – for both the assessment of individual chemical risk and cumulative risk. Also, in reaching a decision about cumulative risk, EPA should consider cumulative exposure to all 21constituents associated with risks that exceed selected benc hmarks, not just the 12 that underwent fullscale analysis.

Dr. William Hopkins, Virginia Polytechnic Institute and State University
Consideration of the combined effects of i ndividual elements within the waste stream. Because this is a complex mixture of wastes, it is inappropriate to treat each constituent separately from one another. Some of the el ements in coal combustion wastes are known to interact with one another, which could l ead to a very different outcome in the risk models. The risk assessment document acknowle dges this shortfall, but it should be corrected in the next draft. At the very least, additivity (which is an underestimate because it does not consider synergistic interactions) should be modeled.

Consideration of multiple exposure routes. Human and ecological receptors may be exposed via diet, water, contact, and/or inhalation and the most likely combinationsDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

45 should be considered in a full scale risk assess ment. This is discussed below (answer to #6) in more detail.

... the risk assessment does not include some of the uncertainties that could be addressed with the data currently available. For exam ple, the risk assessment only considers single exposure pathways. Yet in reality, multiple exposure pathways are important for both humans and wildlife. In many cases the analy tical results are in place for individual scenarios, but the outcomes may be quite different if multiple exposure pathways are simultaneously considered.

A simple scenario would be local citizens w ho use a stream that receives discharge from surface impoundments for recreational purposes (swimming) and sustenance (fishing, trapping, and hunting), while also usi ng groundwater sources for drinking.

EPA Response EPA acknowledges that it did not consider the poten tial cumulative exposure combination of people consuming contaminated fish and drinking water from the same waterbody, because untreated surface water is not considered potable water: surface wate r would be treated by municipal water treatment facilities prior to consumption, reducing contaminan t levels. EPA also did not consider the potential cumulative exposure from contaminated fish and gro undwater in the CCW risk assessment. One reason is that the exposures are likely to occur over different timeframes because of diffe rences in transit time of the contaminant plume to wells versus surface wate rbodies. Also, it may involve different receptors because a resident exposed via groundwater may not be a recreational fisher. Thus, adding risks across pathways would not likely change th e results. This uncertainty is discu ssed in Section 4.4.1 of the risk
assessment.

EPA also acknowledges that some chemicals may inter act in a mixture, causing different health outcomes than would result from individual exposures. However, the quantitative human health benchmarks used by EPA are based on the toxicity of individual chem icals. With only one carcinogen assessed (arsenic), it is not necessary to add carcinogenic risks. Noncarci nogenic risks can be added only for chemicals with toxic effects on the same target organs.

However, the risks for different chemicals peak at different times. To illustrate that chemicals have different arrival times to the modeled wells, peak a rrival times are now presented in Table 4-7 in Section 4.1.1.5 of the risk assessment. Appendix L, which contai ns graphs of the distribution of peak arrival times for several CCW constituents for the groundwater to dri nking water pathway, has also been added to the risk assessment. Because the arrival times tend to diff er between constituents, evaluation of additive risk would be a very complicated endeavor, and the results would be unlikely to change for two reasons. First, because the peak arrival times typically diffe r, there would be instances where a model run would not have two peak (or near peak) risks for constituen ts with the same endpoint. Thus, these additive risks would not always exceed the peaks of individual co nstituents. Second, as pointed out by one reviewer, risks were high enough for single chemicals for human exposure pathways (notably arsenic, which is the only carcinogen). Thus, evaluating additive risks acro ss the different chemicals in each sample would not change the basic conclusions of the risk assessment: that there are potentially significant risks to human health from CCW disposal in landfills and surface im poundments which can be reduced with composite liners. Discussion of this uncertainty has been added to section 4.4.2.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

46 

### 6.2 Toxicity Values

Comment Summary Dr. Vorhees corrected several toxicity values and suggested the use of the IEUBK model for calculating risks to human health from le ad. Additionally, she noted that EPA should include a discussion of the uncertainty associat ed with the toxicity values in the risk
assessment.

Dr. Donna Vorhees, The Science Collaborative Several toxicity values require correction:

1. The PPRTV RfD for aluminum is 1 mg/kg-d, not 2 mg/kg-d (PPRTV).

2. The IRIS RfD for boron is 0.2 mg/kg-d; not 0.09 mg/kg-d as indicated on Table G-1.

3. The IEUBK model should be used to quantify risk for a young child exposed to lead.

4. New July 2008 PPRTV documen tation is available for cobalt:

a. Unit Risk of 2.8E-3 (ug/m3)-1

b. Data insufficient to derive an oral Cancer Slope Factor

c. Reference Dose of 0.02 mg/kg-d (EPA reports this value in the risk assessment)

d. Reference Concentration of 2E-5 mg/m3

For some reason, EPA does not provide PPRTV documentation on-line. Instead, it must be requested by “registered” users. I did not make such a request for all CCW constituents, but EPA should up date the risk assessment to include the latest toxicity values. This should not be labor intensive because the toxicity values were treated as point estimates in the PRA.

EPA’s PRA guidance:

“does not propose probabilistic approaches fo r dose-response in human health assessment and, further, discourages undertaking such ac tivities on a site-by-site basis” (EPA 2001).

I assume that this is why EPA chose not to quantify uncertainty in toxicity values. But EPA should at least discuss uncertainty associat ed with chemical toxicity values in the risk assessment to facilitate inte rpretation of risk results.

EPA Response Toxicity values were chosen based on the Office of Solid Waste and Emergency Response (OSWER) hierarchy (OSWER Directive 9285.7-53). EPA reviewed the toxicity values to confirm their accuracy and determine whether newer values have become ava ilable from EPA or other sources used by EPA for RCRA risk assessments since the CCW risk assessment was conducted in 2002–2003. Results are shown in Table 5 for reference doses (RfDs). As can be seen in the table, toxicity benchmarks have beenDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

47 changed for seven CCW constituents. We have revised th e 50th and 90th percentile risk results to reflect these newer benchmarks, 3 and the results are included in the revised risk assessment.

Table 5. Human Health Benchmark Updates (2002–2009): Reference Doses

Constituent Name CASRN RfD in 2002–2003 Analysis (mg/kg-d) 2002– 2003 Ref a 2009 Updated

RfD (mg/kg-d) 2009 Reference Impact on HQ

Aluminum 7429-90-5 2 A 1 ATSDR chronic MRL (ATSDR, 2009) and PPRTV (U.S. EPA, 2006a) Increase 2x

Barium 7440-39-3 0.07 I 0.2 IRIS (U.S. EPA, 2009a; value added to IRIS July 2005) Decrease 2.9x

Boron 7440-42-8 0.09 I 0.2 IRIS (U.S. EPA, 2009a; value added to IRIS August 2004) Decrease 2.2x

Cobalt 7440-48-4 0.02 SF 0.0003 PPRTV (U.S. EPA, 2008a) Increase 67x Copper 7440-50-8 N/A N/A 0.01 intermediate ATSDR MRL (ATSDR, 2009) no 2002 benchmark Iron 7439-89-6 N/A N/A 0.7 PPRTV (U.S. EPA, 2006b) no 2002 benchmark ATSDR = Agency for Toxic Substances and Disease Registry IRIS = Integrated Risk Information System MRL = minimal risk level PPRTV = provisional peer reviewed toxicity value a 2002–2003 References: A = ATSDR MRLs (ATSDR, 2002) I = IRIS (U.S. EPA, 2002c) SF = Superfund Risk Issue Paper (U.S. EPA, 2001a,b) N/A = not available

EPA decided to use the drinking water maximum cont aminant level (MCL) for lead to estimate risks from drinking water exposure in the CCW risk assessment. Risks from lead exposure were relatively low, well below the risk criterion for landfills and at or sli ghtly above the risk criterion for surface impoundments, and did not drive the risk assessment conclusions. EPA acknowledges that the IEUBK model ( 

### Integrated Exposure Uptake Biokinetic Model for Lead in Children ) may better quantify risk for a young child exposed to lead, as suggested by the commenter. Howe ver, because the MCL was used for nitrate/nitrite as well, EPA has chosen to keep a consistent approach for this particular assessment.

A discussion on uncertainty associated with chemical toxi city values has been added to Section 4.4.3.4 of the revised risk assessment document to facilitate inte rpretation of risk results. This discussion includes a thorough discussion of the hierarchy of toxicity data s ources used to select the best toxicity values for use in the CCW risk assessment. In add ition, Section 3.1.1 and Appendix G contain the toxicity values used for each CCW chemical for screening and the full-scale ri sk assessment. As described in Section 4.4.3.4, the data sources referenced in these sections include thorough discussions of the uncertainty of these values.

3 Risk calculations are linear and so risk results can be directly adjusted to reflect the new benchmarks.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

48 

### 6.3 Target Cancer Risk

Comment Summary Dr. Vorhees questioned why the cancer benchmark of 10 -5 was selected when other EPA programs use a range of 10 -4 to 10 -6 . She suggested that an explanation is necessary.

Dr. Donna Vorhees, The Science Collaborative Why did EPA select a cancer risk benchmark of 10E-5? Some other EPA programs use a range of 10E-4 to 10E-6 range. This topic warrants some discussion; otherwise, how does one explain to a person living on or near a S uperfund site that it is acceptable for them to incur a higher cancer risk than someone liv ing near a WMU used for CCW disposal?

EPA Response The only constituent evaluated for cancer effects is ar senic. In the first footnote of Section 4.0 of the revised risk assessment, EPA notes that:

“The typical cancer risk range used by the O ffice of Solid Waste and Emergency Response is 10 -4
to 10 -6 . In hazardous waste listings, the point of departure for listing a waste is 10 -5

. While the point of departure (10 -5 ) criteria was chosen for use in this risk assessment, EPA reported all arsenic cancer risks above zero. Thus, a risk mana ger could still compare the reported risks with a10 -4 to 10 -6 risk range.


### 6.4 Defining the Exposed Population

Comment Summary Dr. Vorhees questioned the evaluation of only the reasonable maximum exposure scenario. Specifically, she noted that th is does not represent the entire exposure distribution. She suggests that EPA clearly de fine the exposed population of interest in the evaluation based upon the requirements of applicable laws and regulations. Additionally, she notes that the use of exposure data and distributions relevant to the population of interest needs to be demonstrated.

Dr. Donna Vorhees, The Science Collaborative EPA reports that:

“the recreational fisher scenario was used to estimate risks to recreational fishers (and their children) who live near the CCW landfills and surface impoundments and catch and consume fish from a waterbody located adj acent to the [WMU] buffer” (page 2-12).

EPA selected the recreational fisher because:

“they represent the reasonable maximum expo sed individuals and because the streams, lakes, or rivers that are near CCW plants ar e not likely to be used by commercial fishing operations” (page 2-12).

In the context of a PRA, it is illogical to evaluate only the reasonable maximum exposure (RME) scenario because this exposure level represents only part of the entire exposure distribution, albeit the portion of the distribution generally of greatest interest within regulatory contexts. The population that shoul d be the subject of an exposure evaluationDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

49 presumably is dictated by requirements of the applicable laws and regulations, so I do not attempt to define this population. However, EPA must clearly do so. Is this population everyone who lives near WMUs or just recr eational fishers and their families who live near WMUs and also consume fish from water bodies contaminated by WMU constituents? The 90th percentile exposure level will differ depending on how the exposed population is defined. EPA must clearly define the exposed population of interest and demonstrate that it is using expos ure data and distributions relevant to that exposed population (see related comment ary under EPA Condition #5 and Charge Question #4).

EPA Response EPA acknowledges that the recreational fisher scenario was used as a reasonable maximum exposure (RME) scenario for the draft risk assessment (Sections 2.2. 2.2 and 3.8.1), and has clarified that this fisher is assumed to be consuming fish from waterbodi es contaminated by constituents from CCW waste management units (WMUs). Exposure data used for the recreational fisher are described in Section 3.8.2 and Appendix F. In risk assessments used to deve lop regulations under RCRA, EPA traditionally assesses potential risks resulting from an RME scenario in order to ensure that resulting regulation be adequately protective of human health without being excessively c onservative. In this case, data are not available on the populations of recreational fishers living near CCW WMUs and whether or not they are consuming fish from CCW-contaminated waterbodies. EPA believes that the RME assumption is a reasonable and protective alternative given this data gap.


### 6.5 Exposure Factor Assumptions

Comment Summary Dr. Vorhees questioned the use of data from the 1997 Exposure Factors Handbook in the development of intake rate distributions for various exposures, because more current data are available. Additionally, she questioned wh y the variability of fish consumption rates was not considered and suggests that EPA determined rates from the bodies of water affected by WMUs. Minimally, she suggests th at more current fish consumption data could be discussed with regard to the im plications for interpreting risk results.

Dr. Donna Vorhees, The Science Collaborative Drinking Water Intake Rate Distribution. EPA used data from its three-volume exposure factors handbook (1997a-c) to develop many ex posure input distributions. EPA asserts that this handbook “provides the current state of the science concerning exposure assumptions” (page 438). However, exposure studies have been performed since 1997, and the EPA website acknowledges this fact by indicating that the handbook “is being revised to incorporate new data available since 1997” (http://cfpub.epa.gov/ncea/cfm/re cordisplay.cfm?deid=12464).

Fish Consumption Rate. To truly describe the variability in fish consumption rates, EPA needs to identify water bodies affected by WMUs and determine fish consumption rates from these locations. There is no reason to be lieve that data from Maine apply to the whole country, and, as noted above, studies have been conducted and published in the scientific literature since publication of th e exposure factors handbook. Moya (2004) reviews these data, plus EPA will likely fi nd more studies published since this 2004 review. I recognize that all of the cited studi es do not require detailed percentile data that EPA favors, but that is not a reason to i gnore fish consumption rate variability among regions and water body types. At a minimum, EPA could discuss these other fishDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

50 consumption data and implications for inte rpreting risk results (of course, the fish consumption rate distribution first must be included in EPA’s sensitivity analyses).

EPA Response
EPA’s Exposure Factors Handbook (U.S. EPA, 1997a, b, c) represents the current state of knowledge on this topic. EPA’s current guidance on exposure data and use of the Exposure Factors Handbook is consistent with current Agency policy. The Exposure Factors Handbook has been carefully reviewed and evaluated for quality. EPA’s evaluation criteria include d peer review, reproducibility, pertinence to the United States, currency, adequacy of the data collection period, validity of the approach, representativeness of the population, characterization of th e variability, lack of bias in study design, and measurement error. Additional data from the literature published since the Exposure Factors Handbook has not been subjected to the same review by EPA and cannot be included in this analysis at this time.

Distributions for fish consumption data were used in the analysis. Site-specific fish consumption rate data are not available, but the Maine study data, where anglers fished from streams, rivers, and ponds, are consistent with the modeling scenarios used in this ri sk analysis and provide th e detailed percentile data required for a probabilistic analysis. The risk assessme nt report has been revised to discuss the potential impacts of applying the Maine angler consumption rates to the risk assessment, along with uncertainties in applying these rates to other parts of the country where CCW plants are located (see Section 4.4.3.4).


### 6.6 Child/Infant Exposure Assumptions

Comment Summary
Dr. Vorhees commented that no reference is made to EPA guidance regarding the evaluation of exposure during childhood. Additionally, the exposure assessment does not address infant exposure to tap water. She not ed that tap water consumption rates are available for infants. Also, because infants are less likely to consume recreationally caught fish, they should be excluded from the evaluation of this exposure pathway. Dr. Vorhees suggests that data are available fo r tap water consumption rates among children that are more recent than the 1997 Exposure Factors Handbook , and because the newer data set contains more data for infants, it should be used even though it will not likely affect the risk calculations. She noted that the infant age range was excluded from the documentation without explanation. She also suggests that the child fish consumption rates are unnecessarily conservative and other rates should be considered. Additionally she thought it would be useful to include a summary of the more recent literature since the 1997 Exposure Factors Handbook . Dr. Vorhees questions why the maximum value for child exposure duration is 50 years and s uggested that an explanation of how the age distributions are used in modeling is necessary.

Dr. Donna Vorhees, The Science Collaborative The problem formulation discusses the importance of evaluating children separately from adults, given their potentially higher inta ke to body weight ratios. The exposure assessment includes four age groups for individuals younger than 20 to account for variation in exposure over this time interv al [Note: EPA has not documented how it accounted for correlations among exposure across these age groupings. See commentary under EPA Condition #4.] However, the risk assessment includes no references to EPA guidance regarding the evaluation of exposu re during childhood (e.g., EPA 2006a, EPA 2005a), and infant exposure to tap water is not examined. Tap water consumption rates for infants are summarized by EPA (2006a) and discussed under EPA Condition #5.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

51 Infants are less likely to consume recreationall y caught fish; therefore, it is reasonable to exclude this age range from the evaluation of the fish consumption pathway.

EPA’s external review draft of its Ch ild-Specific Exposure Factors Handbook (EPA 2006a) includes some more recent data that are relevant to the evaluation of the drinking water pathway. It summarizes data on tap wa ter intake among children (EPA 2004) that were collected more recently than the tap water consumption rates used in the PRA. These rates were calculated from the combin ed 1994, 1995, 1996, and 1998 Continuing Survey of Food Intakes by Individuals (CSFII), conducted by the United States Department of Agriculture (USDA). In addition to being more recent than the data set in the exposure factor handbook (EPA 2007a), the 2004 data set contains more data for the infant age range (0-1 year old), which was excluded from the PRA without explanation. Including an infant will likely have minima l effect on risk calculations, but exposure during this life stage should be ac knowledged in the risk assessment.

EPA assumed that the child fish consumpti on rate was the same as the adult fish consumption rate. This assumption is un necessarily conservative for many but not necessarily all U.S. subpopulations. EPA shoul d consider reducing the fish consumption rate for a child, using the ratio of adult to child fish consumption rates from various
sources, e.g.,

EPA’s 2002 Estimated Per Ca pita Fish Consumption in the United States; or

West PC, Fly JM, Marans R, Larkin F, Rosenblatt D. 1993. 1991-92 Michigan sport anglers fish consumption study. Prepared by the University of Michigan, School of Natural Resources for the Michigan Department of Natural Resources, Ann Arbor, MI. Technical Report No. 6; May 1993.

It would also be useful to include a su mmary of recent literature beyond what is described in the 1997 exposure factors handbook.

Exposure Duration Distributions. Why is the maximum value for the child exposure duration 50 years? This value makes no sense given the four child age categories that EPA selected. Perhaps if EPA explained how these distributions were used in the modeling, the rationale for this maximum value would be clear. Regardless, a 50-year exposure duration is long but might be appr opriate at some sites, depending on the size and productivity of water bodies near WMUs.

EPA Response EPA acknowledges that infants were not evaluated in the analysis. With respect to drinking water ingestion rates in this analysis, EPA assumed th at most infants are either breastfed or consume commercial formula, and so does not assess potential risk s posed by fish or water consumption by these infants (i.e., modeling start age for child is 1 year). Section 3.8.2 of the risk assessment text has been expanded to include this text.

Dr. Vorhees comments refer to the Child-Specific Exposure Factors Handbook (U.S. EPA, 2008b). The water consumption rates in this document do not differ significantly from those in the 1997 Exposure Factors Handbook and therefore would not be expected to ch ange the results. Exposure duration is based on residence time. For children, the maximum reside nce time is 36 years, based on the study used. Section 3.8.1 of the risk assessment text has been re vised to better explain how the receptors age over the modeled time period:DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

52 “To capture the higher intake-rate-to-body-weight ratio of children, a start age of 1 year was selected for the child receptors. The expos ure duration distribution for cohort 1 (a 1- to 5-year-old) was used to define exposure dur ation for the child receptors for each of the 10,000 iterations in the probabilistic analys is. For each individual iteration, the child receptor was aged through the age cohorts as appropriate until the age corresponding to the selected exposure duration was reached (e.g ., if an exposure duration of 25 years was selected for an iteration, the child was aged fro m 1 year to 25 years, spending 5 years in cohort 1, 6 years in cohort 2, 8 years in cohort 3, and 6 years in cohort 4, for a total of 25 years).”

The risk assessment text in Section 4.4.3.4 has also been revised to include some discussion of the likely impacts the new child exposure factors would have on the results:

“EPA’s child-specific exposure guidance has b een recently finalized (U.S. EPA, 2008b) but was not used in the risk assessment because the wa ter consumption rates a nd body weights provided in the Child-Specific Exposure Factors Handbook (U.S. EPA, 2008b) do not differ significantly from those found in the 1997 Exposure Factors Handbook and would not have changed the results, but the use of the 1997 values may cont ribute some parameter uncertainty. One exception is the distribution of child fish consumption rates used. Here, U.S. EPA (2008b) consumption rates are significantly higher than the 1997 rates u sed in the analysis. This introduces uncertainty into the analysis, and likely underestimates risks in the fish consumption pathway.”

EPA decided to leave the results as originally modeled, but acknowledges the uncerta inty in the child fish ingestion pathway estimates.


### 6.7 Correlations in Input Variables

Comment Summary
Dr. Vorhees commented that correlations we re only discussed in Appendix C as related to how they were accounted for among inputs in the EPACMTP model. No discussion of correlations among other inputs (e.g., fish consumption rate and body weight) was provided.

Dr. Donna Vorhees, The Science Collaborative The authors do not mention the presence or absence of correlations except in appendix C where they discuss how they accounted for correlations among inputs to the EPACMTP model. There is no discussion of correlations among other input distributions, such as fish consumption rate and body wei ght or among age-specific distributions for multiple child age groups that are analyzed in a single M onte Carlo simulation. Only one constituent was modeled at a time; therefore, EPA do es not need to consider correlations among CCW constituent concentrations in drinking wate r or fish. Also, with the exclusion of the dermal exposure pathway, EPA does not need to consider the correlation between body surface area and body weight.

EPA Response EPA acknowledges that human exposure factor data were not correlated. For each modeling run, each exposure factor was selected independently from its own distribution. This introduces some uncertainty because it is possible to select unlikely combinations. In Section 4.4.3.4, EPA has revised the text to discuss this uncertainty and its potential effects on the risk assessment results:DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

53 “As is customary for EPA’s RCRA risk assessments, human exposure factor data were not correlated (i.e., for each modeling run, each exposure factor was selected from its distribution independently), introducing some uncertainty because it is possible to select, for example, a high drinking water rate with a small body weight. However, although a specific modeling run may have had an unrea listic combination of exposure factors, the large number of Monte Carlo iterations performed (10,000) ensures that this is unlikely to significantly affect the risk assessment results.”



### 7.0 COMMENTS RELATED TO ECOLOG ICAL EXPOSURE AND TOXICITY The peer review comments relating to ecological exposure and toxicity are grouped into four subcategories: 1. Cumulative ecological risk 2. Sensitive habitats/sensitive species

3. Indirect effects 4. Eco benchmarks, eco toxicity.


### 7.1 Cumulative Ecological Risk

Comment Summary Dr. Hopkins emphasized the potential importance of cumulative ecological risk, whereby an ecological receptor may be exposed to multiple constituents and/or pathways. For example, amphibians may be subject to both dermal and ingestion exposure.

Dr. William Hopkins, Virginia Polytechnic Institute and State University Ecological scenarios can be even more comple x and consideration needs to be given to the full lifecycle of receptors in the analys is. Take for example amphibians that use surface impoundments for reproductive purposes. Adults can be exposed transdermally because most of the water they acquire cro sses the skin (they don’t typically drink). Concomitantly, they can be exposed trophica lly because the invertebrates they ingest accumulate trace elements from the site. Their eggs are deposited in the water, where developing embryos are directly exposed to dissolved concentrations in the impoundment. Larvae that successfully hatch from these eggs are then exposed to not only the water, but also the contaminated plant material and coal combustion wastes that they ingest directly.

EPA Response EPA acknowledges that cumulative effects can be importa nt for ecological risks. With respect to exposure to multiple constituents, the understanding and character ization of effects to wildlife exposed to multiple constituents concurrently is very limited at this tim e. EPA reviewed some of the literature and the NODA damage case reports for damage cases and provided a summary of the most important in Section 4.2.5 of the revised risk assessment, along with a summary tabl e of ecological impacts noted in the damage cases provided by EPA as part of the NODA. These dama ge cases offer a way to assess this uncertainty, because actual damage cases reflect real world exposur es to multiple CCW constituents. As stated in this section:

“Most of these cases are from surface impoundments and direct discharge into lakes and other water bodies. Along with the published results discussed in Section 4.1.6 , these cases clearlyDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

54 support selenium and arsenic in coal ash as risks to aquatic ecosystems, as well as the adverse impacts of coal ash on terrestrial vegetation.”

Regardless of the importance of cumulative effects, single-exposure ecological risks were sufficiently above the risk criteria that it is clear that CCW di sposal poses potentially significant threats to ecological health without considering cumulative ecological exposures.


### 7.2 Sensitive Habitats/Sensitive Species

Comment Summary Dr. Hopkins commented that the risk assessment does not consider sensitive habitats or species. Rather, this is listed as an uncertainty to be dealt with on a site-specific basis. Dr. Hopkins requested additional discussion of this issue, including “how it can be dealt with on a site specific basis” and “specific steps to be taken to determine what habitats and species deserve special consideration.”

Dr. William Hopkins, Virginia Polytechnic Institute and State University The only other major comment I have relates to the exclusion of sensitive habitats and sensitive species from the analysis. This is lis ted as an uncertainty that will need to be dealt with on site-specific basis. I agree with this conclusion, but it would be useful to see this idea receive more prominence in the c onclusions of the document. How can it be dealt with on a site-specific basis? What speci fic steps should be taken to determine what habitats and species deserve special consid eration (besides the obvious threatened and endangered species)?

EPA Response An expanded discussion of sensitive habitats and species is provided in the scenario uncertainty section of the revised risk assessment (Section 4.4.1). In gene ral, the presence of sensitive habitats and species is inherently a site-specific issue that was not eval uated in the CCW risk assessment. Nevertheless, the benchmarks used were conserva tive, and only the lowest (most conservative) benchmark for each constituent and pathway was used. Thus, the conditions represented by simulations in the upper end of the risk distribution (higher risk scenarios) may reas onably characterize many situations with sensitive species/habitats.



### 7.3 Indirect Effects

Comment Summary Dr. Hopkins emphasized the potential importance of indirect ecological effects (e.g., depletion of food resources), suggesting that exclusion of indirect effects from the risk assessment underestimates risk. He requested at a minimum discussion of indirect effects as a major uncertainty. Dr. Harvey requested additional discussion of “how ecosystems can be adversely affected by discharges of solutes in unexpected ways” and how this issue is uncertain due to insufficient scientific understanding.

Dr. William Hopkins, Virginia Polytechnic Institute and State University
Consideration of indirect effects of toxic constituents in coal combustion wastes on ecological systems. Currently, the risk a ssessment is based on the assumption that all effects of contaminants are direct (i.e., direct toxicity). In reality, when contaminants are discharged into ecological systems, they ha ve community wide-effects that are often mediated through ecological interactions. I ndirect effects include responses such asDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

55 depletion of food resources, alteration of pr edator-prey interactions, and modified hostpathogen interactions. It is well known that th e indirect effects of contaminants can be just as important, and in some cases more important, than direct effects. This has been demonstrated for a wide variety of organic pollutants and metals (e.g., Fleeger, et al., 2003). Similarly, a variety of ecological effects have been documented from coal combustion wastes that could have broa d, indirect effects on other community constituents (e.g., Reviewed in National Academy of Sciences Report, pages 89-90). Failure to consider indirect effects in the m odels (at the very least, as a major uncertainty) underestimates risk.

Dr. Charles Harvey, Massachusetts Institute of Technology I recognize that the hazard quotient approach may be the only workable method for attempting to quantify ecological risks. However, I think the report would be strengthened by a candid discussion of how aqua tic ecosystems can be adversely affected by discharges of solutes in unexpected ways . The state of the science is simply not sufficient to predict how different species, le t alone ecosystems, will respond to an input of a mixture of solutes. Many examples exist of how aquatic environments are impacted in surprising ways by shifts in water chemistry.

EPA Response EPA has expanded the report discussion on indirect eco logical effects and the associated uncertainty in Section 4.4.2 of the revised risk assessment. Howeve r, the comment raises an important issue regarding the scale of potential effects and, in particular, the ecological significance of those effects. For any given facility, the spatial scale of potential contamination would affect a very small proportion of the home range for typical species; the determination of the im pact on food supply and habitat quality with regard to the landscape and overall health of the animal s is unknowable in a national-level assessment (and difficult to understand in the majority of site-sp ecific assessments). In addition, many species are opportunistic feeders and will seek other areas if food s ources decline, regardless of whether the source of the stress to the food supply is anthropogenic or natural in origin.


### 7.4 Ecological Benchmarks, Ecological Toxicity

Comment Summary Dr. Basta was concerned that the ecological benchmarks used in the risk assessment may have been overly conservative and that many are at or below background soil concentrations, mentioning lead in particular (specifically, that the 95th percentile of the natural background of lead in soils in the United States is 38.8 mg/kg [Smith et al., 2005], higher than the value of 28 mg/kg used in the risk assessment). He suggested EPA Ecological Soil Screening levels (Eco-SSLs) as an alternative. In the example of lead, the Eco-SSL of 1,700 mg/kg is 60 times higher than the value used in the risk assessment.

Dr. Nicholas Basta, Ohio State University The ecological database used to generate ecological benchmarks and calculate ecological HQs is overly conservative, and at times, unr ealistic. Many of the ecological benchmarks are at or below background soil concentrati ons. Ecotoxicological data were identified from summary documents prepared at the Oa k Ridge National Laboratory (Efroymson et al., 1997a,b).

Many of the Soil Criteria Ecological Benchmarks in the Draft are well below natural background levels in soil. Natural background so il levels of the constituents considered inDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

56 Table H-3 of the Draft Ecological risk assessment are not phytotoxic. Contaminant soil concentrations must be much greater than natural background concentrations to observe ecotoxicity at Draft Benchmark effects levels. For example, the Soil Criteria of 28 mg/kg Pb was used in the risk assessment (Table H- 3) for soil invertebrates as the terrestrial receptor. However, the 95th percentile of the natural background of Pb in soils in the United States is 38.8 mg/kg (Smith, 2005). The Soil

Criteria is below background soil Pb levels . We do not observe phytotoxicity to soil invertebrates at 28 mg/kg Pb in natural soils.

USEPA has established Ecological Soil Screeni ng Levels (Eco-SSLs) of contaminants in soil that are protective of ecological recept ors that commonly come into contact withand/or consume biota that live in or on the soil. The derivation procedures for the EcoSSLs represent the collaborative effort of a mu lti-stakeholder team consisting of federal, state, consulting, industry, and academic par ticipants led by the USEPA, Office of Solid Waste and Emergency Response. The Eco-SSL values were derived separately for four groups of ecological receptors: plants, soil invertebrates, birds, and mammals. These values are presumed to provide adequate protection of terrestrial ecosystems. Eco-SSLs are derived to be protective of the conser vative end of the exposure and effects species distribution and are intended to be applied at the screening stage of an ecological risk assessment. The detailed procedures used to derive Eco-SSL values are described in separate documentation (USEPA, 2003).

The USEPA Pb Eco-SSL for soil invertebrat es is 1,700 mg/kg Pb (USEPA, 2005). Thus, the Eco-SSL of 1,700 mg/kg is considered to be protective of the conservative end of the exposure and effects species distribution. The Eco-SSL is 60 times higher than the Soil Criteria Ecological Benchmark of 28 mg/kg used in the Draft. Why the difference? The Eco-SSL values are based on an exhaustive review of previously published peerreviewed literature from soil invertebrate studies that included natural soils and soil invertebrate toxicity endpoints (USEPA, 2003). Because of its scope, Eco-SSL values are better indicators for ecotoxicity screening than the study used to establish Ecological Benchmarks in the Draft.

The Ecological Benchmarks in Table H-3 should be compared with (a) the 95th percentile concentration of natural soils for each constituent, and (b) the USEPA EcoSSL values. The Soil Criteria should be reeval uated and justified if the Benchmark is below natural soil concentrations (Smith, 200 5) where we know toxicity does not occur. At a minimum, a table comparing the three should be reported in the Draft.

EPA Response EPA agrees with Dr. Basta’s comments that the most recent ecological data, such as the Ecological Soil Screening levels (Eco-SSLs), should be used when a ppropriate. When the ecological benchmarks used in the CCW risk assessment were developed, EPA used a very different methodology to establish benchmarks for the soil community relative to the Eco-SSL effort. The CCW methodology was intentionally conservative and was based on species sens itivity distributions that include a wide range of soil fauna. In the 10 years since the CCW ecological benchmarks were developed, there has been a considerable effort by EPA and others to improve the data and methods used in benchmark derivation.

EPA compared the CCW risk assessment ecological benchmarks with the Eco-SSLs and as might be expected from the 10 years since the CCW chemical st ressor concentration limits (CSCLs) were derived, there are differences. Many values are somewhat lower (e.g., the Eco- SSLs), especially when the lowestDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

57 value among multiple receptors are used. For example, the lead Eco-SSL of 1,700 ppm mentioned by Dr. Basta is for soil organisms. Although this is considerab ly higher than the 28 mg/kg value used in the risk assessment, the lowest Eco-SSL for lead is 11 mg/kg for birds, which woul d be selected for use in the risk assessment according to current EPA practice to use the lowest of mu ltiple benchmarks when evaluating ecological risk.

Dr. Basta’s suggestion that the benchmark used for lead is below background is somewhat misleading. The Smith et al. (2005) paper indicates that the background levels for lead range from 5.3–319 mg/kg, with a mean value of 23.3 mg/kg (these data reflect two transects across the country). The expression of toxicity with regards to soil biota (as well as plants a nd other flora and fauna) is inherently a site-specific phenomenon; areas with high natural concentrations of lead tend to be colonized by lead-tolerant species and have lead in more stable, less bioavailable forms than lead from CCW disposal. In areas with relatively low background concentrations of lead, the introduction of higher lead levels may result in a significant change in the soil community that could adve rsely affect the ecosystem. In addition, the lowest Eco-SSLs for lead, 11 mg/kg for birds and 56 mg/kg for mammals, bracket the value used in the CCW risk assessment (28 mg/kg), so the current risk assessment results for lead are within the range that would be expected if the results were updated to the ne wer values and are reasonably protective of the most sensitive species.



### 8.0 COMMENTS ABOUT THE MONT E CARLO ANALYSIS APPROACH The peer review comments relating to the Monte Carlo analysis approach are grouped into four subcategories: 1. Use of conservative parameter estimates 2. Variability versus uncertainty 3. The calculation of percentile probab ilities as opposed to an overall probability

4. Comments related to the numerical stability of Monte Carlo results 5. Impacts on groundwater as a resource.



### 8.1 Use of Conservative Parameter Estimates

Comment Summary Dr. Harvey expressed concern about the use of conservative estimates for some parameters. He emphasized that the ideal approach would be to estimate the actual risk and associated uncertainty rather than wei ghting the results conservatively. “From a decision making point of view, it is much clean er if the distribution of risks represents a best estimate. Then attitudes towards risk (i .e., being “protective”) can be considered when using this distribution to make a decision.”

Dr. Charles Harvey, Massachusetts Institute of Technology The logic of this assessment also skates on thin ice when “conser vative” or “protective” values of parameters are selected. The idea of a Monte Carlo model is to generate a best estimate of the actually probability distribution. While there is a clear motivation to use protective values of a parameter when assessi ng environmental contaminants, choosing a value that gives an upper bound on risk, rather than the best estimate of risk, may be an easy way out of the difficult job of assessing probabilities and this approach may muddle the overall assessment. It is often easier to say, “I know that this parameter can’t possibly be beyond the value A” than to support the claim that a parameter has a value B with a certain standard deviation. The problem with this approach is that it moves risk aversionDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

58 into the characterization of the distribution of outcomes. From a decision making point of view, it is much cleaner if the distribution of risks represents a best estimate. Then attitudes towards risk (i.e. being “protectiv e”) can be considered when using this distribution to make a decision. (Note: for the cancer risk from arsenic the distribution may be best collapsed into a single pr obability, as discussed in above in C.)

EPA Response EPA agrees in the ideal of producing best estimates of actual risk, and most input data distributions were developed to satisfy this objective. These distributio ns were developed from av ailable data and included aquifer data, soil type, WMU data, climate data, distance to groundwater wells, distance to surface waterbodies, constituent concentrations, water flow data, and many of the human exposure factors. However, in some cases, input data were not fully av ailable to characterize the input distribution and its uncertainties. In such cases, values th at might be considered conservativ e were selected for the draft risk assessment. Thus, further discussion of these uncertainties has been added to Section 4.4.3.


### 8.2 Variability versus Uncertainty

Comment Summary Dr. Vorhees pointed out that the risk assess ment does not formally separate variability from uncertainty, which makes it challenging to identify opportunities to reduce uncertainty.

Dr. Donna Vorhees, The Science Collaborative The authors discuss variability and uncer tainty. Distinguishing between them quantitatively is important for determining o pportunities to reduce uncertainty in risk estimates. EPA’s input distributions are in tended to represent national variability. In addition EPA reports that “Uncertainty also was considered in the analysis by using reasonable ranges and distributions when variables were not known exactly. For example, when a soil texture or groundwater flow c onditions could not be precisely assigned at a site, multiple soil types or hydrogeologic environments would be sampled based on the soil and aquifer types that are likely to be pr esent at the site.” (page 4-27) As a result, EPA has not separated uncertainty and vari ability quantitatively but has instead blended them in some input distributions. To the extent possible, EPA guidance recommends separating them to help in identifying opportunities to reduce uncertainty.

Simulation inputs were not defined in a manne r that would allow for a clear quantitative distinction between the contribution of va riability and uncertainty to the simulation output. This limitation of the Monte Carlo simulations is discussed below under Condition #5.

EPA Response EPA acknowledges the ideal of separating variability from uncertainty when po ssible in a probabilistic risk assessment, and that EPA did not do so here. Because the resources required to acquire the necessary data on facilities burning coal and managing the wast e onsite are not commensurate with the expected improvements in the CCW risk assessment results, EPA w ill not pursue such analysis in the revised risk
assessment.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

59 

### 8.3 Calculation of Percentile Probabiliti es as Opposed to an Overall Probability

Comment Summary Dr. Harvey stated that the distributions of cancer risk estimated in the risk assessment can be reduced to a single probability and that this result can provide particular insight into the results.

Dr. Charles Harvey, Massachusetts Institute of Technology For the cancer risks from arsenic, the Monte Carlo model generates probability distributions of cancer probabilities. A probability distribution of binary probabilities can be reduced to a single probability. For example, consider the case where you are adverse to tails in a coin toss. If there is a 0.6 proba bility that the coin is fair and 0.4 probability that the coin is weighted so that it lands tails with 0.75 probability, then the probability of getting tails is: 0.6*0.5 + 0.4*0.75 = 0.6. What you really care about is the 0.6 chance of tails, and the distribution of probabilities used to calculate this risk is of little value to you. Likewise, for the cancer risks one could si mply calculate the final cancer risk (the probability of cancer) over the Monte Carlo generated distribution of risks. This calculated probability may be the most important estimate, more useful than the 0.5 and 0.9 quantiles of probability provided in this report. The assessment should certainly report this probability, and it would also be useful if the report discussed how distributions of probabilities should be cons idered. For the hazard quotients, the full distribution is necessary because hazard quotients are not probabilities. A clear explanation of the use of estimated distri butions would help the reader understand the conditions for which distributions are useful, and the conditions under which they may be converted to a single probability without losing important information.

EPA Response EPA acknowledges that such a simplification may be possible. However, EPA policy states that decisionmakers should be given a range of risk estimates from at least two points on a distribution of risk estimates; a central tendency estimate, and a “high-end” estimate (U.S. EPA, 1992.). This is consistent with EPA policy, and provides more useful information to the decision-makers.


### 8.4 The Numerical Stability of Monte Carlo Results

Comment Summary Dr. Vorhees requested an evaluation of the st ability of the Monte Carlo results. In other words, is the number of simulations suffici ent, so that the results do not change significantly with more simulations?

Dr. Donna Vorhees, The Science Collaborative EPA performed Monte Carlo simulations of 10,000 iterations per waste type/WMU type/constituent combination. However, they do not provide quantitative evidence of numerical stability of 50th and 90th percentile estimates using this number of iterations. This information needs to be in the report. In addition, when EPA reports results by liner type, e.g., Table 4-1, it is reporting results from fewer than 10,000 iterations, with the number of iterations per liner type a functi on of how frequently WMUs with each liner type were selected in the Monte Carlo simulation.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

60 EPA Response Figure 2 shows average risk versus th e number of simulations for arsen ic under selected CCW disposal scenarios, with the green and pink lines depicting la ndfill risks for ash and ash mixed with coal mill rejects, and the blue and red lines depicting surface impoundment risks for the same two waste types. In each case, the averages of the model results converge as the number of simulations increases. Consistent with other probabilistic risk analyses conducted by EPA, around three to five thousand simulations are necessary for convergence in most case s, and the 10,000 simulations perfo rmed for this analysis are more than sufficient for all modeled cases.

Figure 2. Example showing convergence of Monte Carlo simulations of up to 10,000 realizations.


### 8.5 Impacts on Groundwater as a Resource

Comment Summary Dr. Harvey requested a discussion of the potential for groundwater contamination to impact the availability of groundwater as a resource in the future. Similarly, Dr. Rowe questioned whether groundwater use patterns w ould change in the future, thus affecting the potentially exposed population (e.g., highe r population density leading to more wells closer to the source). Dr. Rowe requested an evaluation of the sensitivity of the results to the well placement assumptions.

Dr. Charles Harvey, Massachusetts Institute of Technology The report would also benefit from a discussion of the non-health and non-ecological costs of polluting groundwater: namely, the loss of groundwater as a resource to futureDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

61 generations. Even with no adverse health impact s, there is an important adverse effect of contaminating groundwater. The groundwater downgradient of a WMU may be lost as a useable source of water. This adverse effect of WMUs could be assessed by characterizing the volume of groundwater lost as a potential resource near WMUs. The volume of this lost resource is substantially la rger than the volume of groundwater that is actually contaminated because the pattern of groundwater contamination downstream of a source is nearly impossible to completely ch aracterize, so the groundwater over a large area will not be used.

Completely ignoring health effects because of the belief that no one will drink contaminated water, and shifting the analys is to purely protection of water resources, would be an extreme approach, and I would not support such a limited analysis. However, I think that the report would benefit from a candid discussion of the actual effects on the population near a WMU, a discussion of the possible numbers of people who would really drink contaminated well wa ter and the detrimental effects of loosing valuable water supplies.

Dr. Kerry Rowe, Queens University In the absence of an actual survey of wa ter well usage around existing CCW facilities, I am not aware of a better source of data than that available for MSW landfills. EPA recognized this as “one of the larger uncer tainties in this analysis” (RA, pg. 4-37) however the statement that “EPA believes th at the MSW well distance distribution used is protective for CCW landfills and surface im poundments” (RA, pg. 4-37) appears to be nothing more than that – a belief – with no ju stification being given for this article of faith. Also, as noted in response to Question #5 , the analysis extends for 10,000 years but no consideration appears to have been gi ven to likelihood of increased demand for the use of groundwater in the next 30-100 years (let alone the next 10,000 years) which presumably means more wells and makes the pr otection of groundwater likely even more important for the future than it is now. It would have been good to have examined the implications of wells being closer than th at indicated by the MSW survey to assess how sensitive the RA is to that assumption (one can infer that it will be important – but not how important in terms of risk).

EPA Response EPA acknowledges Dr. Harvey’s comments with regard to future groundwater resources and has added language on this topic to the scenario uncertainty section of the revised report (Section 4.4.1):

“The risk assessment did not explicitly consid er potential impacts on the availability of groundwater in the future (e.g., contamin ated groundwater becoming unsuitable for consumption), but the results do clearly indicat e that there can be a reduction in resource availability if CCW is improperly disposed. However, the scope of the risk assessment was to evaluate human health and ecological effects associated with current waste disposal practices and conditions, and a qua ntitative evaluation of potential future reductions in groundwater availability as a consequence of CCW disposal practices was not conducted as part of this analysis.”

Regarding Dr. Rowe’s comment on receptor well loca tion, EPA agrees that there was not a better data source available and that distance to receptor wells may have an influence on the results. However, it is possible that the distribution of dist ances to receptor wells used is simila r to that of actual CCW sites. The remainder of Dr. Rowe’s comment has been addressed in Section 3.1 of this comment response document.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

62 

### 9.0 COMMENTS RELATED TO THE DO CUMENTATION AND MISCELLANEOUS COMMENTS The peer review comments relating to documen tation are grouped into seven subcategories: 1. Insufficient documentation to reproduce the analysis independently 2. Graphical input/output distributions, truncation values 3. Documentation of sensitivity analysis 4. Discussion of general risk-susceptibility patterns (e.g., climate-specific risk)

5. Discussion of worst-case scenarios 6. Discussion of sulfate, phosphate, and chloride 7. Background concentrations.



### 9.1 Insufficient Documentation to Re produce the Analysis Independently

Comment Summary Dr. Basta, Dr. Rowe, and Dr. Vorhees stated that the documentation is incomplete and that an independent reviewer could not reproduce the analysis. For example, Kd values generated by MINTEQA2 used for EPACMTP are not presented in the documentation.

Dr. Nicholas Basta, Ohio State University The parameterization of the risk models is clear but the inputs and outputs of several modeling components are absent and must be provided. A thorough review of the CCW risk assessment was not possible because input data used or data generated for the risk assessment was not provided. The inputs and outputs for each step of the risk assessment should be provided. For example, Kd values generated by using MINTEQA2 (used as input data for EPACMTP) are critical but ar e not reported. The Kd value selected can greatly influence the outcome of the risk assessment. Review of Kd values, and other intermediate modeled outputs , is critical. The reviewer cannot determine whether Kd values used were “reasonable.” Similarly, modeled groundwater concentration used as input data for the risk assessment should be provided for review. In short, the modeling methodology should be more transparent.

The parameterization of the risk models is clear, but the inputs and outputs of several modeling components are absent and must be provided. The inputs and outputs for each step of the risk assessment should be prov ided. In short, the modeling methodology should be more transparent.

Dr. Kerry Rowe, Queens University The documentation of the parameterization is at a variable but generally high level. It is adequate to give a general idea of how the RA was conducted and, except for the limitations as noted in response to other questions, appears reasonable at the level provided. However the level of detail is insu fficient for a reviewer to comment in detail on the distributions that were input to the Monte Carlo simulations and hence it is also not possible to reproduce the RA based on the documentation provided for review. Thus, except at a very high level, it is not possible to comment on the completeness or reasonableness of the distributions that were input to the Monte Carlo simulations. This is a significant short-coming if the objective of the documentation is to allow a complete review of the approach that was adopted in arriving at the results. That said, it is alsoDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

63 recognized, that full documentation will likely be quite voluminous and would take much longer to review than has been allocated.

Dr. Donna Vorhees, The Science Collaborative My general comment is that the risk assess ment does not provide sufficient information to allow results of the analysis to be independently reproduced. My more detailed comments follow.

EPA describes the foundation of the Monte Ca rlo simulation as the source data derived from the EPRI survey, explaining that these da ta were combined with data from EPA’s CCW constituent database to conduct a si mulation of 10,000 iterations per waste type/WMU type/constituent combination. EPA also should provide the name of any software program and/or annotated computer code that it used to perform these simulations.

EPA Response EPA acknowledges that the documentation of the inputs and intermediate outputs could have been more transparent. However, EPA notes that many of th e inputs EPA used are directly discernable from the appendices. Constituent concentrations and leachate va lues are presented in Appendix A. Facility sizes, liner types, ash types, and locations are presented in Appendix B. Distances to wells, distances to surface waterbodies, soil data, climate data, and hydrological data are presented in Appendix C. To improve transparency, EPA has added Table D-1 in Appendix D to show 

### selected percentiles of Kd sampled from the MINTEQA2 isotherms for every waste management modeling scenario conducted in the CCW risk assessment for the groundwater pathway . Again, including all of the MINTEQA2 inputs and outputs would be too voluminous to increase the transparency of the document. EPA refers the commenters to the more complete documentation available in other reports that have previously been peer reviewed and released by EPA. For example, the MI NTEQA2 approach and results are documented in the EPACMTP Technical Background Document (U.S. EPA , 2003b). EPA acknowledges that the additions and discussions of inputs in the document are not suff icient for duplication of the results, and can provide the input and output files in Microsoft Access format upon request.


### 9.2 Graphical Input/Output Dist ributions, Truncation Values

Comment Summary Dr. Vorhees requested graphical presentation of input and output distributions. She also requested further explanation of the minimum and maximum truncating values.

Dr. Donna Vorhees, The Science Collaborative EPA provides much of the documentation requi red for a PRA, including exposure input distributions, models used to predict surface water, fish, and groundwater concentrations, and models used to predict noncancer h azard quotients and cancer risks. EPA discusses many of the most important assumptions inhere nt in their choices of input distributions. However, EPA does not include input distributions and output distributions in graphical format, which would aid interpretation and use of results, particularly given the relatively large differences between some corresponding 50th and 90th percentile results. It would be useful to see plots that include the input distributions and the data used to construct the distributions as a supplement to goodness-of-f it statistical test results. EPA does not explain why it chose the minimum and maximum values for truncating input distributions or whether truncation affected parameter estimates for the distributions.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

64 EPA explains that HQs for the child resident were consistently higher than HQs for the adult resident; therefore, only child HQs are pr esented. This is a reasonable choice, but it would be helpful to define which of the four childhood age groups corresponded to the child resident and to label tables accordingly.

EPA Response As noted above, EPA can make the input and output files available upon request.



### 9.3 Documentation of Sensitivity Analysis

Comment Summary Dr. Vorhees requested more complete document ation of the sensitivity analysis. She also stated that she could not find the report that originally presented the sensitivity analysis (U.S. EPA, 2005).

Dr. Donna Vorhees, The Science Collaborative EPA briefly summarizes the sensitivity analyses that it describes in another report (EPA 2005b). I was unable to find this report onlin e. The sensitivity analyses are an integral part of the PRA, vital to proper interpre tation and use of results, and should not be reported in a separate document.

EPA’s summary is not sufficient to understa nd the full scope of the sensitivity analysis effort, the procedures that EPA used, and exte nt to which EPA was able to distinguish influential sources of variability from influe ntial sources of uncertainty. EPA excluded human exposure factors from the sensitivity analysis, describing them as sources of variability that could not be changed or mana ged in anyway. Nevertheless, if one of these inputs strongly influenced risk results, this might indicate that EPA needs to take additional steps to determine how well its input distributions accurately reflect variability on a national level. The fish consumption rate distribution is based on data from only one region, so it will be particularly important to determine its relative influence on risk estimates.

EPA Response EPA acknowledges the omission of the original sens itivity analysis from the docket. EPA has since updated the sensitivity analysis (U.S. EPA, 2009c) so that it clearly describes the methodology and results summarized in Section 4.3 of the risk assessment. The revised sensitivity analysis will be placed in the docket along with the revised risk assessment.


### 9.4 Discussion of General Risk-Susceptibility Patterns such as ClimateSpecific Risk

Comment Summary Dr. Rowe and Dr. Harvey requested an analys is of risk susceptibility patterns—i.e., what conditions lead to more or less relative ri sk. Dr. Rowe focused on potential climate patterns (e.g., less risk in dry conditions).

Dr. Charles Harvey, Massachusetts Institute of Technology This report misses the opportunity to describe which WMU are the most dangerous. It would be useful to know the conditions wher e the risks are the highest, and then to mapDRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

65 where these locations are most likely to exis t. This could be accomplished simply by taking the top, say 20%, of realizations w ith the worst risk and analyzing what these realizations have in common. Some characteris tics will be trivial, such as a short distance to the well or water body, but others, such as the geochemical and hydrologic parameters, might provide useful insight into where high-quality liners are particularly important.

Dr. Kerry Rowe, Queens University While it is understood that the objective was to do a RA for the entire US using the distributions of sites in the ERPI (1997) report, it needs to be recognized that (a) climate and (b) composite-liner design and constr uction have a profound impact on the WMU infiltration rate for landfills (correctly identif ied as a key parameter affecting the RA, e.g. pg. 4-40). These are discussed below.

The potential risk of a disposal facility in an arid climate may be substantially different to those in a humid climate and hence the nature of, for example, the level of liner that is needed to meet risk criteria may vary from one climatic region to another. The current RA may underestimate risk for humid zones and overestimate risk in arid zones. Thus for the purposes of developing management op tions for CCW (Objective 2), more cost effective management strategies may be deve loped by considering what is required for different climatic regions rather than adopting a one-size-fits-all approach.

EPA Response EPA refers Dr. Harvey to Section 4.3 of the risk assessment, where a summary of the sensitivity analysis is provided. EPA finds that across constituents and path ways, the most sensitive drivers of risk appear to be leachate concentration and infiltration rate (i.e., liner type). Other parameters that appear to be sensitive for specific constituents and pathways in clude aquifer hydraulic conductivity and gradient, adsorption isotherm coefficient, depth to groundwat er, receptor well distance, and waterbody flow rate. Although EPA did not analyze the top 20% of mode l runs as the commenter suggests, the sensitivity analysis results confirm that higher input values for th ese variables are likely to lead to higher risk results.

With respect to the difference between humid and arid climates, the sensitivity analysis suggests that the risk results were not as sensitive to these cl imate variables as the commenter suggests.



### 9.5 Discussion of Worst-Case Scenarios

Comment Summary Dr. Rowe emphasized that worst-case scenario s may exceed the risks modeled in the risk
assessment.
Dr. Kerry Rowe, Queens University The discussion of receptor populations is consid ered reasonable given the broad scope of the study. However it should be acknowledged that there may be special cases where the assessment may underestimate risk. Under th ese circumstances, site specific studies would be warranted. These situations are un likely to occur ofte n but the possibility should be recognized in the development of any regulations based on the RA. For example, the consumption of fish assumes recr eational fishing and this is, in most cases, likely to be reasonable. However if the recep tor water body was used by subsistence fish consumers (native harvesters as identified in th e comments by Dr. Foran) then the risks could be substantially higher than modeled.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

66 EPA Response EPA acknowledges that a range of outcomes may occu r. The draft risk assessment only showed the results for the 50th and 90th percentiles. Neverthele ss, EPA appreciates the importance of site-specific analysis. Therefore, the risk asse ssment has been expanded to include discussion of actual damage case scenarios (in Sections 4.1.6 and 4.2.5) that may be closer to the “worst-case” scenarios discussed by Dr. Rowe than the reported results from the risk assessment.


### 9.6 Discussion of Sulfate, Phosphate, and Chloride

Comment Summary Dr. Rowe questioned why sulfate, phosphate, and chloride were excluded from the analysis. He requested more justification.

Dr. Kerry Rowe, Queens University Although it is acknowledged that some of the contaminants not considered (e.g. sulfate, phosphate and chloride; RA pg. 2-3) could “ pose an ecological hazard if concentrations are high enough”, there is inadequate discussion of the justification of this exclusion and its rationalization.

EPA Response Sulfate, phosphate, and chloride were excluded from the CCW risk assessment because human health and ecological benchmarks that are compatible with th e constituent-specific approa ch used in the risk assessment are not available for these constituents . Available damage case data reviewed and incorporated in Section 4.1.5 in the revised risk assessment document demonstrate that there potential impacts of such CCW constituents on drinking wate r, as shown by exceedances of secondary MCLs or state standards. However, EPA modeled only the constitu ents most likely to result in human health or ecological risks, and because secondary MCLs ar e not health-based, EPA did not consider these constituents in the analysis.


### 9.7 Background Concentrations

Comment Summary Dr. Vorhees requested discussion of background concentrations of constituents in CCW.

Dr. Donna Vorhees, The Science Collaborative EPA did not discuss results in light of bac kground concentrations of CCW constituents in the absence of a release from WMUs. Su ch discussion would improve the risk
assessment.
EPA Response EPA acknowledges that a discussion of background concen trations of CCW constituents would be helpful to a risk manager. However, the scope of the analysis was specifically to analyze the potential impacts of releases associated with CCW management. Alt hough background concentrations certainly may contribute to risk when present, or can be higher than some of the media concentrations calculated in the CCW risk assessment, EPA generally considers only th e incremental increase in risk associated with constituents in wastes when conduc ting risk assessments under RCRA.DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009

67 

### 10.0 REFERENCES

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