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-09-004 

 DRAFT 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 Modeling 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 Availability of the Vanderbilt Study .............................................. 19

4.2 2006 DOE Facility Data and Liner Scenarios ........................................................................... 23

4.3 Alternative Waste Disposal/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 Exposure 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/Sensitive 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 Parameter 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 ...................................................................... 59
 DRAFT 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 Documentation 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 Patterns 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 ................................................................................................................................. 67
 DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009 

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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 information 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 assessment to evaluate the potential exposure to CCW in 

landfills and surface impoundments. 

As stated in the NODA, the risk assessment was independently 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 used in geochemical modeling 

▪ Dr. Charles Harvey, Massachusetts Institute of Technology: use of probabilistic approaches 

used in the analysis in the context of fate and transport of contaminants in the environment 

▪ Dr. William Hopkins, Virginia Polytechnic 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 (Docket 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 Constitution 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 

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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 timeframe. 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 risk over a 10,000-year 

period. Shifting land use, geological and climate 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 concentration in the leachate data remains 

constant throughout the 10,000-year modeling period is not realistic. Leaching of 

contaminants will likely occur in much shorter time periods and decrease after reaching a 

maximum. The peak contaminant concentration 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 contaminant 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 change 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 horizon should be further clarified in the revised 

risk assessment. The choice of a maximum 10,000-year time horizon does not mean that all simulations 
 DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009 

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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 concentrations at the receptor point return to zero after the 

concentration peak or for the maximum simulation time of 10,000 years, whichever comes first.” 

Although groundwater concentration predictions are simulated after the peak or maximum average 

concentration is observed, the post-peak or post-maximum 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, respectively. 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 significantly, 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 leachate 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 Transformation 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 otherwise 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 support a different approach. Thus, for this draft risk 

assessment, the leaching duration was assumed to be finite (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 patterns 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 scenario 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 contaminant concentrations in the leachate remain 

constant throughout the 10,000-year modeling 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 total waste concentration available to be 

leached, and it was assumed that 100% of the constituent in the waste could leach from 

the landfill. The nonlinear fate and transport solution used for metallic constituents in the 

unsaturated zone module of 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 of U.S. EPA, 2003b). Although a leaching 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 

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The adoption of a simplified leaching profile to support a non-linear sorption approach in 

the unsaturated zone offered a greater benefit and defensibility to the overall approach 

than assuming linear partitioning and a depleting 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 determined 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 value 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 the ability of the MINTEQA2 model-generated 

Kd values used in EPACMTP to determine realistic 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 properties critical to chemical processes values. 

This could be conducted using column studies 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 are lumped into one value, Kd, calculated by 

MINTEQA2. This is a common modeling approach, although this grossly oversimplifies 

and is likely not an accurate approach for several CCW chemical constituents that are 

controlled by precipitation-dissolution processes in soil that cannot be accurately 

modeled using Kd values. 

Chemical processes are mainly 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 leachate and capable of being transported. The 

contaminant is not transported if it is not dissolved. 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 arsenite). Soils vary greatly in the type and 

amount of these sorbents. The MINTEQA2 model is very limited in its ability to model 

sorption. MINTEQA2’s database is limited to three sorbents: Fe sorbent as the mineral 
 DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009 

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goethite, Fe sorbent as hydrous ferric oxide, and organic matter. Fe sorbent as goethite 

and organic matter were selected to calculate nonlinear isotherms and subsequent Kd 

values in the Draft. The choice of Fe sorbent is important because goethite is a much 

poorer adsorbent than hydrous ferric oxide and 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 calculated 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 approach used by MINTEQA2 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 important sorption of hydrous Fe, Al, and Mn 

oxides soil minerals were not reported. 

Arsenic solubility and transport in soil is greatly 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 will 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 transport in soils with moderate to high levels 

of hydrous Fe, Al, and Mn oxides could be orders of magnitude. 

However, some insight into impact of exclusion of hydrous Fe, Al, and Mn oxides on As 

in leachate is provided from the Draft: “Sensitivity analysis results indicate that for most 

of the scenarios evaluated (over 75 percent), the risk assessment model was most 

sensitive to parameters related to groundwater flow and transport: WMU infiltration rate, 

leachate concentration, and aquifer hydraulic 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 adsorbed 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 scenarios studied?” is comparing Kd used for 

transport modeling using EPACMTP in the Draft with Kd values for select soils from 

CCW facilities. Kd values for As and other select contaminants should be determined for 
 DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009 

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a set of soils with a range of soil chemical properties critical to chemical processes 

values. This could be conducted using column studies 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. 

Arsenic speciation was considered in the Draft 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 arsenic(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 the result that drinking water standards are 

rarely exceeded in offsite groundwater in spite 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 concentrations) 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 environmental 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 results 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 much 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 overestimated 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 single 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 that 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 the 
 DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009 

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modeled source of organic matter and its reactivity; Geochemists who are interested in 

redox transformations may criticize the lack 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 words, our ability to model reactive transport 

may not be sufficiently developed to construct good models over a wide range of 

conditions without constraining 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 sites 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. Second, 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 isotherms, 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 very 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 plume 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 solute 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 modeling would be more mechanistic than using linear 

and nonlinear partitioning coefficients; however, such modeling is not a practical approach for a 

nationwide analysis because (1) the data collection 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/precipitation processes is a widely known and accepted method 

in the scientific literature and the groundwater modeling community, provided the values of Kd used are 

appropriate. 

Nevertheless, we have added discussion to the risk assessment 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 modeling 
 DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009 

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(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 individual values in the Kd isotherms (depending on metal 

concentrations, and master variables such as pH, organic carbon, and iron oxide concentrations). Second, 

there is limited information generally available about the conditions under which published values were 

measured. This comparison would be very time consuming 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 assessment 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 periphery 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 mixes 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 assumption 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 species were selected specifically for their respective 

sorption isotherm behavior; Kds derived from As(III) isotherms tend to decrease as pH increases (which 

is typical of most metal species examined in the 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

Full Mix

No Mix Full Mix × − % Change =

where 

CNo Mix = Simulated peak receptor well concentration 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 

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CFull Mix = Simulated peak receptor well concentration for a select percentile based on a fully 

mixed assumption (mg/L) 

Table 1 shows the percent change in peak receptor well concentrations assuming no mixing occurs for a 

range of percentiles. 

Table 1. Change in Peak Receptor Well Concentrations 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 species indicate that As(V) has a sensitivity to pH that leads to increased 

receptor well concentrations under the no mixing assumption (i.e., when the leachate pH is used to 

determine Kd in the saturated zone) relative to the 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 associated 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. Basta 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 model runs with higher risk typically involved shorter 

distance to receptors. Thus, one would expect less time and distance for adsorption and precipitation to 

occur. Updated discussion of the uncertainties with arsenic and selenium speciation is reported in Section 

4.4.2 of the revised risk assessment. 

Finally, with respect to Dr. Harvey’s comment, the contaminant transport formulations in both the 

unsaturated and saturated modules of EPACMTP correctly include a sorption term that represents the 

transfer of dissolved contaminants from the liquid phase to the solid phase by way of a retardation factor 

or an isotherm describing the change in sorbed concentration with respect to dissolved concentration. The 
 DRAFT 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 Comparisons 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 confidence 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 effort expended to conduct a nationwide 

assessment of risks posed by CCW disposal practices. 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-truthing.” 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 assessment 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 performed 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 the 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 measured 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 complexity and uncertainty that perhaps 

could be reduced if EPA examined a representative 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) and 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 submitted to the Massachusetts Department of 

Environmental Protection. Public commentary submitted from industry and government 

(State of Wisconsin) representatives noted the availability of additional data, and these 

leads should be explored to determine if high 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’s 
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11

risk assessment embodies an enormous amount of work and creative, thoughtful 

modeling. But, at a minimum, EPA needs to provide a compelling reason for not 

attempting some site-specific analyses, which is the most effective way to reduce 

uncertainty in the PRA. 

EPA reports that it has data for antimony, cobalt, 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 limits, 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 transport 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 collection 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 Results vs. Reported Groundwater Exceedences 

Damage Casesb

Constituent 

2007 Risk 

Assessmenta

Human 

Health 

Effectsc

Cosmetic/ 

Aesthetic 

Effectsd

State 

Standarde

Consistent 

Results as of 

2007f

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

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Damage Casesb

Constituent 

2007 Risk 

Assessmenta

Human 

Health 

Effectsc

Cosmetic/ 

Aesthetic 

Effectsd

State 

Standarde

Consistent 

Results as of 

2007f

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, the 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 showed 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 risk 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 constituents are those for which the risk assessment and the damage cases agree. 

Some constituents in this group showed risk in the model and also appeared in the damage cases (arsenic, 

boron, cadmium, lead, molybdenum, and selenium). Other constituents in this group did not show the 

potential for risks in excess of the risk criteria from the 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 which the risk assessment and the damage cases did not 

agree. Four modeled constituents (antimony, cobalt, thallium, 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 these constituents, (2) not enough time has passed to see the remaining 

constituents appear in damage cases, (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, nickel, and vanadium). Because all that is known is 

that these constituents have the potential to pose a risk to human health, they 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 cases 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 damage 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 consistent 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 exist at proven damage case sites. This inconsistency 

could indicate that (1) the risk assessment was conservative with respect 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 

Assessmenta Damage Casesb

Consistent Results as of 

2007c

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, the 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 contaminated, 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. Harvey’s comments focused on the assumption 

that the wells used in this assessment are contaminated (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 which the assessment handles uncontaminated 

wells, plume characteristics, groundwater–surface water interactions and vertical 

contaminant concentration across a screened interval 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 curves 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 are 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 presents puzzling conceptual problems stemming 

from the fact that the calculated health risks are independent of the probability that the 

contaminated well actually exists. The following example illustrates the difficulty. In 

comment II.B. below, I will discuss how, wherever 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 the framework used in 

this report, the calculated health risks would 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 pathway must generate a well position, short 

plumes will actually have worse adverse health effects because the wells generated by the 

Monte Carlo study will be closer to the source 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 examples. 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 that 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. After 
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15

reading the public comments, I believe that this discussion would be valuable --it appears 

that the authors of these public comments may not have understood this issue. 

Public understanding of the dangers of contaminated 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. Because studies such as this have documented 

the serious health dangers of contaminated groundwater, and because severe 

contamination has often been found down-gradient 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 the 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 distribution of distances measured in 1988 for 

WMUs. This framework is basically reasonable. 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 important 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 the 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 directly 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 such 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 dispersivity. 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 direction than in the direction of flow, and hence 
 DRAFT 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 spreading on contamination of wells differs from 

horizontal spreading because wells are often screened 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 thickness of the aquifer. Thus, the vertical 

concentrations might be better approximated by averaging concentrations vertically. 

Finally, geologic formations are often layered 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 often 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 water 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 uncertainties in this analysis” (RA, pg. 4-37) 

however the statement that “EPA believes that the MSW well distance distribution used 

is protective for CCW landfills and surface impoundments” (RA, pg. 4-37) appears to be 

nothing more than that – a belief – with no justification 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 given 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 protection 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 that indicated by the MSW survey to assess how 
 DRAFT 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 assessment used a distribution of well distances 

measured at municipal solid waste (MSW) landfills. Is this distribution appropriate for 

CCW 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 say 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 avoided 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 will 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 drinking water wells as surrogate measures for 

the distance between WMUs and drinking water wells. EPA appropriately cites the 

uncertainty associated with this surrogate information 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. This 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 cases 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 that 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 explanation 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 across the contaminant plume centerline and 

perpendicular to the groundwater flow direction. 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 assumed 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 saturated 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 the 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 transfer 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 the groundwater-surface water interface.” 

The comments regarding the placement of receptor wells 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 explanation 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 lateral 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 practice, 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 the actual level of exposure. This possibility exists 

because simulated concentrations are point observations, not concentrations from simulated pumping 

wells that may capture both contaminated and pristine groundwater, depending on factors identified in Dr. 

Harvey’s comment. Nevertheless, EPA opted to use the current approach, which may conservatively 

overestimate the receptor well concentrations. 

Dr. Harvey’s comment on dispersivity partially highlights the issue of transparency. The documentation 

for this risk assessment did not include comprehensive tables detailing modeling input parameters, their 

values or distributional characteristics, and the source of the data used; however, these values are, in 

many cases, provided in the EPACMTP Background and 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 where 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 dispersion 
 DRAFT 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 additional 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 coal 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 toxicity characteristic leaching procedure (TCLP) 

data and the availability of more representative 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 (HELP) 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 surface impoundment source concentrations 

11. Modeling surface impoundments only before closure 

12. Modeling landfills only after closure. 

4.1 Use of TCLP Data and the Availability of the Vanderbilt Study 

Comment Summary 

This subcategory addresses comments from Dr. Basta, Dr. Hopkins, Dr. Rowe, and Dr. 

Vorhees. These reviewers expressed 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 assessment, 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 municipal 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 results.” Therefore, additional information that improves the 

waste leachate data set would be expected to significantly reduce key uncertainties. 

Selection of appropriate analyte data for landfills 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 concentration 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 concentration from CCW. Improved leachate methods, 

designed for CCW, should be used to generate data in the Risk Assessment. 

TCLP was developed to characterize waste behavior 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 calculations 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 laboratory 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 Assessment 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 used 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 comparison 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 actual leachate data would have to be 

extended to include all the input data 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 the 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 leaching 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 are not adequate for characterizing how coal 

combustion wastes will behave in a field setting. There are a number of serious problems 

with using this test to predict long-term behavior 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 best available for the time being, but when 

better alternatives are developed (there are 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 TCLP 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 parameters used in the RA are reasonable for 

arsenic and or potentially unconservative for selenium. 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 regarding 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 analyzes 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 chemistry and leaching behavior (EPA 2006b). EPA 

explains that these data were not included because the risk assessment was performed in 
 DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009 

22

2003 (page 1-1). EPA acknowledges that inclusion 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 representative 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 currently 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 statement raises another question: how will the 

results of this risk assessment be used? Are they 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 more recently (See Figure 4-5). EPA argues that the 

data sets are similar but this figure suggests otherwise. 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 equilibration 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 developed 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 Enhanced Sorbents for Mercury Control (U.S. EPA, 2006c). Although the 

2006 report was focused on mercury, it also addressed arsenic and selenium. The risk assessment 

included a comparison and discussion of all the data used in the risk assessment versus the 

Vanderbilt/EPA data for these three elements. This discussion can be seen in Section 4.4.3.1. Specifically, 

EPA notes that: 

“...the TCLP data do appear to encompass the range of variability in CCW leachate 

concentrations that have been measured in more recent studies. [...] For arsenic, the CCW values 

bracket the range found in the other studies. Selenium 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 in 

Characterization 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 aluminum, antimony, arsenic, barium, boron, cadmium, 

chromium, cobalt, lead, molybdenum, selenium, and thallium. EPA does not intend to conduct a new risk 
 DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009 

23

assessment based on the eight facilities in this new study because such a small number of sites would not 

be as representative of a nationwide distribution as the 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 assessment 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 identified 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 associated with the facilities that existed in 

1995 but the data does need to be carefully interpreted if it is to be used to assess the risk 

of future facilities and for developing future 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 the 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 acknowledge the issues raised above in response to 

Question #1. 

As discussed in response to Question #1 (and 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 existed 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 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 the 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 development 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 evaluation of risks for “clay-lined” is only done 

for sites that are clay-lined). This is appropriate for evaluating the risk associated with 

these facilities, however even better insight regarding the effect of the three liner 

conditions (“unlined”, “clay-lined” and “composite-lined” ) could have been obtained by 

examining all three liner possibilities for all 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 that 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 substantially 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 development 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 evaluation of risks for “clay-lined” is only done 

for sites that are clay-lined). This is appropriate for evaluating the risk associated with 

these facilities, however even better insight regarding the effect of the three liner 

conditions (“unlined”, “clay-lined” and “composite-lined” ) could have been obtained by 

examining all three liner possibilities for all 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 that 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 gained by installing/mandating different liners. 

EPA Response 

Because they reflect liner conditions in 1995 (the time 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 assessment. 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 risks 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 the risk assessment considered alternative 

CCW disposal scenarios. Specifically, codisposal with wastes generating more acidic 

conditions may significantly impact the mobility of metals. Dr. Hopkins suggested that 

the risk assessment should consider CCW beneficial reuse options such as structural fill. 

Dr. Kerry Rowe, Queens University 

However as previously indicated in response 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 than 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 and reuse, is not considered in the risk assessment. While 

it would clearly be far beyond the scope of this 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”. The 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 using 
 DRAFT Response to Peer Review Comments on the CCW Risk Assessment – September 1, 2009 

26

appropriate safeguards such as liners. Similarly, use of these wastes as a structural fill, 

particularly in large monofills without safeguards, 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 codisposal 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, including 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, 2006c, 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 source 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 was 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 Tetra 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 sites 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 construction 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 from 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 conclusion 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 considerable research suggesting that (a) 

geomembranes have a finite service life and (b) leakage through composite liners 

depends significantly on both the nature of the clay component of the composite liner 

(CCL or GCL) and the presence of wrinkles when 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 factors (see response to Questions #1 above) 

would improve the RA and increase the calculated 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 amount of water related to consolidation can 

be shown to be small. Also, composite liners with a compacted clay liner (CCL) can be 

expected to leak much more than those with GCLs because of the poorer interface 

transmissivity (many orders of magnitude higher) for CCLs. This approach adopted 

introduces an unconservative bias to the corresponding 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 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 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 water content and density etc.), (b) construction 

quality control and assurance during construction, 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 constructed 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 rater 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-feet 
 DRAFT 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. erring on the high side for clay-lined sites). 

As previously noted above, the leakage calculated as described in the RA (pg. 3-20) is 

likely to underestimate the leakage through composite liners and hence underestimate the 

risk associated with these liners. That said, properly constructed composite liners are still 

going to substantially reduce the risks associated with CCW disposal compared to 

unlined and clay-lined sites and this refinement will not change the general conclusions 

of the RA. 

... as discussed in response to earlier questions above, while variability was considered 

with respect to MWU characteristics, there does not appear to have been adequate 

consideration of uncertainty associated with composite-lined sites. 

EPA Response 

The composite liner leakage rates used for this risk 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. Several literature references discuss 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 following simple calculation illustrates that this effect 

is realistic at least in some cases. Assuming the porosity 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 consolidation (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 attributable to consolidation has not been quantified 

accurately to date. Due to the absence of an approach for estimating consolidation flow rates versus 

leakage, EPA excluded clay liner data from the estimated distribution of composite liner leakage rates. 

Although there is clear literature support for the assumed 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 not 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 rates versus 6% for the original risk assessment 

distribution). In addition, the DAFs are significantly smaller for the alternative leakage rates. 

Although changing the liner leakage assumptions would change the risk assessment results for compositelined facilities, it would not alter the general conclusion about the relative risk of different liner scenarios 

(composite lined landfills pose significantly less potential risk than unlined landfills). This latter point 

was emphasized by Dr. Rowe in his review of the 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

DAF

Percentile

CCW RA leakage rates

Alternative leakage rates

Figure 1. DAF Comparison for different leakage rates. 

With respect to the clay liners themselves, the 2009 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 differently 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 impoundments 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 focused on the assessment’s use of half of 

the detection limit when a constituent’s concentration was not detected. He felt that lower 

values should have been used, especially for arsenic 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. Guide 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: Supporting 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 assess the risks while acknowledging the shortcomings (the 

likely overly conservative nature) of such an analysis. 

Dr. Nicholas Basta, Ohio State University 

The “Calculation of average constituent concentrations 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 disposal 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 reported “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 critical because As drives the Human Health 

drinking water exposure pathway standard, and it is very possible that Hg would have 

driven the Ecological Risk exposure pathway. The analytical laboratory instrumental 

detection limits are not reported. Cursory inspection of the data in Appendix A-2 suggests 

improved analytical methods should have been used to achieve lower detection levels that 

would have decreased or eliminated the large 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 that concentration may increase in the future 

(RA, pg. 4-34) is also questionable. At the time the study was conducted, an alternative to 

ignoring mercury would have been to do analyses 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 the 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 samples 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 generate the upper-percentile risk levels. Therefore, from a 

statistical standpoint, the replacement of these nondetects with half the detection limit or some other 

suggested value will not change the 90th percentile risk levels generated by the risk assessment. 
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31

Although the treatment of nondetects could change the 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 limit. Because this is one of the most sensitive inputs, 

such a change would likely significantly raise the 50th 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 overestimate of risk. Alternatively, setting all nondetects 

to zero would be even more inappropriate. Given the 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 where 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 extreme 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 Guidance 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 for surface impoundments) is very different. Here, 

virtually all of the measurements were nondetects. For this reason, mercury was eliminated from the 

analysis to eliminate the potential (and unknown) bias. Thallium and antimony were likewise eliminated 

from analysis in the surface impoundment scenarios. Clearly, the risk assessment with respect to these 

constituents would be enhanced by new waste measurements 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 issue of mercury, thallium, and antimony is discussed 

in Section 4.4.3.1 of the revised risk assessment. 

4.6 Effect of Waste 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 constituents from the site. However, compaction 

technologies reduce water penetration through the materials, rendering them less likely to 

produce leachate. If the data are available, this 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 potential compaction of CCW waste over time. Such 

compaction could decrease the hydraulic conductivity and the associated water infiltration. 

However, no readily available data were identified to support an analysis of the influence of 

CCW compaction on infiltration rates. The current approach would tend to overestimate 
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32

infiltration rates compared to a model that would adjust the hydraulic conductivity over time due 

to compaction. EPA believes this is an appropriately conservative assumption given the lack of 

the information needed to accurately model 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 landfills has been the subject of some recent 

criticism in the literature however the limitations of this model, and the potential 

implications of its use, have not been adequately addressed in the document. 

Any assessment of infiltration through a landfill 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 the 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 fundamental 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 over many years. This is well discussed in the HELP 

model documentation, as well as the EPACMTP Parameter/Data Background Document (U.S. EPA, 

2003a) as discussed and referenced in Section 3.4.2.1 of the revised risk assessment. 

4.8 Scenarios Where the Water Table Is Above the Disposal Unit 

Comment Summary 

Dr. Rowe expressed concern about the disconnect 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 below 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 about the modeling only considering disposal 

sites where the bottom of the facility is above the water table while claiming that CCW 

landfills are “often constructed in mines and 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 enforced to control the head in the landfill 

to below that outside the landfill (hydraulic containment), then the risk of these sites will 

indeed be higher than modeled. However given that the general conclusions from the 
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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 handle the scenario where the water table is above 

the disposal unit, and has added a statement to model 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 disposal units do extend below the water table. 

Because waste intersecting the saturated zone 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 conclusions 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 Capture 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 will 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 provide 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 the 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 reconsider the risks from a very different waste 

stream. Importantly, the amount of Hg that will 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. 
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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 acknowledgement 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 completed 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, not potential future changes 

associated with emissions controls. EPA added a statement 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, and 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 mercury 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 mercury leachability in CCW can vary depending on 

the flue gas mercury controls used at a power plant.” 

4.10 Use of Porewater Data for Surface Impoundment Source Concentrations 

Comment Summary 

Dr. Rowe expressed concern about estimating surface impoundment source 

concentrations using measured concentrations in pore water collected from underlying 

sediments. The pore water concentrations are likely 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 geochemical 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 variability but will have a 

bias to lower values. Thus while the modeling does consider retardation mechanism as 

contaminants migrate through the soil, the use of a source term that is artificially low 
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35

because of sorption/precipitation that occurred 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) downstream from impoundments (i.e., direct surface 

water contamination, independent 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 combustion wastes are needed to assess risk to 

animals (deer, dabbling ducks, grazing amphibian larvae, certain invertebrates) that 

directly ingest the solid waste materials when 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 vertebrates [especially birds and mammals]) or deposit 

their eggs directly in the water (amphibians, 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 applied to represent a range of site-specific 

scenarios. In addition to calculating resultant concentrations after 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 estimating 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 represent the leachate seeping from the bottom of the 

impoundment than would bulk surface impoundment waste concentrations. The pore water is in 

direct contact with the waste, so these concentrations 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-term 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 behaves more as a landfill than a surface impoundment. 
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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 leachate into the underlying soil with greater 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 revised risk assessment, including a discussion 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 there 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 are much higher during surface impoundment operation 

because the higher hydraulic head in an operating impoundment drives wastewater into the 

underlying soil with greater force than infiltration 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, especially in unlined units. Thus, even if the 

post-closure period were modeled, the corresponding results would not be as high as the peak 

annual leachate concentrations used in the analysis.” 

In addition, EPA notes that the arrival times for 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 arrival time for peak boron and selenium VI risk is 

74 years. Thus, even with further leaching occurring 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 operating landfills should have been considered, 

because the greatest risks may occur during site operations when the waste products are 

directly exposed to wind and rain. 
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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 model used in the CCW risk assessment did not 

model leaching from landfills before closure, as the waste 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 scenarios 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 drinking, 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 impoundments requires that the wastes be mixed 

with water before being pumped into the impoundments. Overall, the findings of the risk 

assessment are consistent with this. That is, risks are higher for surface impoundments 

than for landfills. However, the risk assessment grossly underestimates risks from surface 

impoundments to humans and the environment because it fails to consider the two most 

important exposure pathways for invertebrates, fish, wildlife, and humans. For surface 

impoundments, the risk assessment only recognizes leaching as a release mechanism and 

1) groundwater-to-drinking water and 2) groundwater-to-surface water as exposure 

pathways. This is incorrect. While leaching into groundwater is clearly a concern at these 

sites, the predominant exposure pathways relate to 3) direct exposure to the slurried 

wastes in the impoundments and to 4) direct exposure to the effluent from the 

impoundments discharged offsite. I highlight the problems with overlooking these routes 

of exposure below. 
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38

Direct Exposure Onsite. Plants and animals 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 water, and reproduction. Examples include game 

(e.g., deer, migratory waterfowl, and raccoons) and non-game species (migratory 

passerine birds, wading birds, amphibians, and 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 contain large quantities of fine, suspended 

particulate matter (e.g., fly ash) and high concentrations of dissolved trace elements 

resulting from the slurry. Several important groups 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., frogs 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 come in contact with the water and coal 

combustion wastes in and around surface impoundments, but is especially problematic 

for species with highly permeable skin (e.g., amphibians). 

From an ecological perspective, all of the groups of organisms mentioned above are 

important. It is possible that such sites act as population sinks in the landscape for species 

such as amphibians and certain birds (e.g., green 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 exposures to both surface impoundment wastewater (using 

data from the 1998 risk assessment) and surface impoundment porewater were considered and compared. 

Although the porewater risk estimates were intended to represent surface waters contaminated with 

groundwater, they were based on undiluted porewater concentrations, and thus could represent direct 

exposure to wildlife as well. They also include trophic exposure increases (i.e., bioaccumulation up the 

food chain). Results from the 1998 CCW risk assessment, which were described in the 2002 CCW risk 

screening document, support the conclusion that ecological risks from direct exposure to surface 

impoundment wastewater are significant, which agrees with the commenter. These conclusions can be 

further supported with damage incidents in published literature. As a result, EPA did not evaluate these 

risks further in the full-scale assessment, considering the screening results and damage cases to be 

sufficient evidence that these risks are significant. EPA has integrated 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 cases 
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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 species 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 species are particularly concerning. For example, 

a study on landfilled coal combustion wastes with 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 exposure, it is less likely that hunters (and 

their families) could be protected from the hazards of eating waterfowl that have used 

surface impoundments. Migratory waterfowl often use surface impoundments and in 

some cases (depending on where the impoundment 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 mechanism 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 wastes, 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 risk assessment and as a result, the assessment 

grossly underestimates exposure and resultant risks. 

EPA Response 

We agree that this human exposure scenario was not addressed in either the screening analysis or the fullscale CCW risk assessment, and that risks may be significant if humans consume wildlife contaminated 

by CCW constituents. The scenario uncertainty section of the risk assessment (Section 4.4.1) has been 

revised to add a clear statement that this pathway was not considered: 

“In addition, not all possible exposure pathways were evaluated. For example, the risk assessment 

did not consider potential indirect exposure to humans through game species that may have been 

exposed to surface impoundment waste (e.g., deer drinking surface impoundment water). This 

represents a potential uncertainty in the analysis.” 
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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 downstream 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 animals can experience direct exposure to coal 

combustion wastes downstream from surface impoundments. Depending on how the site 

is engineered, surface water in impoundments 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 dilution factors are large and offsite transport is 

typical. While concentrations of trace elements 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 resuspension of coal combustion wastes and the direct 

discharge of suspended particulates into downstream habitats. These problems can be 

further exacerbated in large rivers that receive discharge from surface impoundments at 

multiple power plants, creating cumulative exposure 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 artificial selection for more antibiotic-resistant 

communities 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 sites. However, recreational use of impacted 

streams and rivers could also result in significant contact with water and sediments, 

creating multiple exposure pathways. For fish and wildlife, particular attention should be 

devoted to species that utilize areas that favor deposition of trace elements and fine 

particulates, including areas of reduced flow (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 risk assessment. Direct exposure in surface 

impoundments and downstream from surface impoundments 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) waste disposal risk assessments do not address direct 

discharges from impoundments to surface waters because 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 the 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 for the management of RCRA waste disposal, 

EPA recognizes that CCW surface impoundment effluent may pose additional risks” 

5.4 Exposure During Landfill Operation (Before Closure) 

Comment Summary 

Dr. Hopkins pointed out that the risk assessment 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 and resulting exposure 

pathways associated with surface impoundments (see comments [Major Criticism] above 

regarding exposure to contaminants in surface 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 the wastes during site operations will result in 

increased risks for workers and residents in the areas adjacent to transport routes and 

landfills. Although this concern 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 decades and this period may represent the greatest 
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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 does 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. However, 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 through 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 receptors 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 exceedance (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 plants 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 impacts of CCW on amphibians at DOE’s Savannah River 

Plant (Hopkins et al., 2006). Section 4.2.5 of the final 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 the 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 pathways 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 why exposure pathways were eliminated. I 
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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/index.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 for 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 made available in the docket at www.regulations.gov 

(Document #: EPA-HQ-RCRA-2006-0796-0470). The screening 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 HUMAN EXPOSURE AND TOXICITY 

The peer review comments relating to human exposure and toxicity are grouped into seven subcategories: 

1. Combined chemical effects, additive 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 identified by a review of EPA’s existing data 

related to relative proportion of CCW constituents at WMU (waste management unit) 

locations as well as predictions of CCW constituent concentrations in drinking and 

surface water near WMU locations. Additionally, she noted that cumulative risk needs to 

be reported from the perspective of the receptor and that cumulative exposure should be 

considered for all 21 constituents with risks exceeding chosen benchmark values. 

Dr. Hopkins noted that it is not appropriate to treat each constituent 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. Additionally, he notes that multiple exposure 
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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 fish from the water body and drinking water 

from groundwater (without considering interaction). It is not clear why the potential for 

people both consuming fish and drinking water 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 exposure 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 tackled 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, cumulative 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 exposure to which no one is exposed in reality. 

Specifically, the 90th percentile exposure level 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 predictions 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 cause 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 benchmarks, not just the 12 that underwent fullscale analysis. 

Dr. William Hopkins, Virginia Polytechnic Institute and State University 

Consideration of the combined effects of individual 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 elements in coal combustion wastes are known 

to interact with one another, which could lead to a very different outcome in the risk 

models. The risk assessment document acknowledges 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 combinations 
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45

should be considered in a full scale risk assessment. 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 example, 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 analytical 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 who use a stream that receives discharge from 

surface impoundments for recreational purposes (swimming) and sustenance (fishing, 

trapping, and hunting), while also using groundwater sources for drinking. 

EPA Response 

EPA acknowledges that it did not consider the potential 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 water would be treated by municipal water treatment 

facilities prior to consumption, reducing contaminant levels. EPA also did not consider the potential 

cumulative exposure from contaminated fish and groundwater in the CCW risk assessment. One reason is 

that the exposures are likely to occur over different timeframes because of differences in transit time of 

the contaminant plume to wells versus surface waterbodies. 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 the results. This uncertainty is discussed in Section 4.4.1 of the risk 

assessment. 

EPA also acknowledges that some chemicals may interact 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 chemicals. With only one carcinogen assessed (arsenic), it 

is not necessary to add carcinogenic risks. Noncarcinogenic 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 arrival times are now presented in Table 4-7 in Section 

4.1.1.5 of the risk assessment. Appendix L, which contains graphs of the distribution of peak arrival times 

for several CCW constituents for the groundwater to drinking water pathway, has also been added to the 

risk assessment. Because the arrival times tend to differ 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 differ, there would be instances where a model run would 

not have two peak (or near peak) risks for constituents with the same endpoint. Thus, these additive risks 

would not always exceed the peaks of individual constituents. 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 across 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 impoundments which can be reduced with composite 

liners. Discussion of this uncertainty has been added to section 4.4.2. 
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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 lead. Additionally, she noted that EPA should 

include a discussion of the uncertainty associated 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 documentation 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 update 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 for dose-response in human health assessment 

and, further, discourages undertaking such activities 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 associated with chemical toxicity values in the 

risk assessment to facilitate interpretation 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 available 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 been 
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47

changed for seven CCW constituents. We have revised the 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 

Refa

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 contaminant 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 slightly 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. However, 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 toxicity values has been added to Section 4.4.3.4 of 

the revised risk assessment document to facilitate interpretation of risk results. This discussion includes a 

thorough discussion of the hierarchy of toxicity data sources used to select the best toxicity values for use 

in the CCW risk assessment. In addition, Section 3.1.1 and Appendix G contain the toxicity values used 

for each CCW chemical for screening and the full-scale risk 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. 
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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 Superfund site that it is acceptable for them to 

incur a higher cancer risk than someone living near a WMU used for CCW disposal? 

EPA Response 

The only constituent evaluated for cancer effects is arsenic. In the first footnote of Section 4.0 of the 

revised risk assessment, EPA notes that: 

“The typical cancer risk range used by the Office 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 manager 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 this does not represent the entire exposure 

distribution. She suggests that EPA clearly define 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 adjacent to the [WMU] buffer” (page 2-12). 

EPA selected the recreational fisher because: 

“they represent the reasonable maximum exposed individuals and because the streams, 

lakes, or rivers that are near CCW plants are 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 should be the subject of an exposure evaluation 
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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 recreational 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 exposure data and distributions relevant to that 

exposed population (see related commentary 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 waterbodies 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 develop 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 conservative. 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 why 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 that more current fish consumption data 

could be discussed with regard to the implications 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 exposure 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/recordisplay.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 believe 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 the exposure factors handbook. Moya (2004) 

reviews these data, plus EPA will likely find more studies published since this 2004 

review. I recognize that all of the cited studies do not require detailed percentile data that 

EPA favors, but that is not a reason to ignore fish consumption rate variability among 

regions and water body types. At a minimum, EPA could discuss these other fish 
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50

consumption data and implications for interpreting 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 included 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 the 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 risk analysis and provide the detailed percentile data 

required for a probabilistic analysis. The risk assessment 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 noted 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 for 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 suggested 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 intake 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 interval [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 exposure 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. 
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51

Infants are less likely to consume recreationally 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 Child-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 water 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 combined 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 minimal effect on risk calculations, but exposure 

during this life stage should be acknowledged in the risk assessment. 

EPA assumed that the child fish consumption rate was the same as the adult fish 

consumption rate. This assumption is unnecessarily conservative for many but not 

necessarily all U.S. subpopulations. EPA should 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 Capita 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 summary 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 appropriate 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 that most infants are either breastfed or consume 

commercial formula, and so does not assess potential risks 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 change the results. Exposure duration is based 

on residence time. For children, the maximum residence time is 36 years, based on the study used. 

Section 3.8.1 of the risk assessment text has been revised to better explain how the receptors age over the 

modeled time period: 
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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 exposure duration distribution for cohort 1 (a 1- 

to 5-year-old) was used to define exposure duration for the child receptors for each of the 

10,000 iterations in the probabilistic analysis. 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 from 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 been recently finalized (U.S. EPA, 2008b) but was 

not used in the risk assessment because the water consumption rates and 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 contribute 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 used 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 uncertainty in the child fish 

ingestion pathway estimates. 

6.7 Correlations in Input Variables 

Comment Summary 

Dr. Vorhees commented that correlations were 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 weight or among age-specific distributions for multiple child 

age groups that are analyzed in a single Monte Carlo simulation. Only one constituent 

was modeled at a time; therefore, EPA does not need to consider correlations among 

CCW constituent concentrations in drinking water 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: 
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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 unrealistic 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 ECOLOGICAL 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 complex and consideration needs to be given to 

the full lifecycle of receptors in the analysis. Take for example amphibians that use 

surface impoundments for reproductive purposes. Adults can be exposed transdermally 

because most of the water they acquire crosses the skin (they don’t typically drink). 

Concomitantly, they can be exposed trophically 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 important for ecological risks. With respect to exposure 

to multiple constituents, the understanding and characterization of effects to wildlife exposed to multiple 

constituents concurrently is very limited at this time. 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 table of ecological impacts noted in the damage cases 

provided by EPA as part of the NODA. These damage cases offer a way to assess this uncertainty, 

because actual damage cases reflect real world exposures 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 clearly 
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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 disposal 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 listed 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 conclusions of the document. How can it be 

dealt with on a site-specific basis? What specific steps should be taken to determine what 

habitats and species deserve special consideration (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 general, the presence of sensitive habitats and species is 

inherently a site-specific issue that was not evaluated in the CCW risk assessment. Nevertheless, the 

benchmarks used were conservative, 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 reasonably 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 assessment 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 have community wide-effects that are often 

mediated through ecological interactions. Indirect effects include responses such as 
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55

depletion of food resources, alteration of predator-prey interactions, and modified hostpathogen interactions. It is well known that the 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 broad, 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 models (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 aquatic 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, let 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 ecological effects and the associated uncertainty in 

Section 4.4.2 of the revised risk assessment. However, 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 impact on food supply and habitat quality with regard 

to the landscape and overall health of the animals is unknowable in a national-level assessment (and 

difficult to understand in the majority of site-specific assessments). In addition, many species are 

opportunistic feeders and will seek other areas if food sources 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, unrealistic. Many of the ecological benchmarks 

are at or below background soil concentrations. Ecotoxicological data were identified 

from summary documents prepared at the Oak 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 soil levels of the constituents considered in 
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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 Screening Levels (Eco-SSLs) of contaminants in 

soil that are protective of ecological receptors that commonly come into contact with 

and/or consume biota that live in or on the soil. The derivation procedures for the EcoSSLs represent the collaborative effort of a multi-stakeholder team consisting of federal, 

state, consulting, industry, and academic participants 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 conservative 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 invertebrates 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 reevaluated and justified if the Benchmark is 

below natural soil concentrations (Smith, 2005) 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 appropriate. 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 sensitivity 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 stressor concentration limits (CSCLs) were derived, 

there are differences. Many values are somewhat lower (e.g., the Eco-SSLs), especially when the lowest 
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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 considerably 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 would be selected for use in the risk 

assessment according to current EPA practice to use the lowest of multiple 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 and 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 adversely 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 newer values and are reasonably protective of the most 

sensitive species. 

8.0 COMMENTS ABOUT THE MONTE 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 probabilities 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 weighting the results conservatively. “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 “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 “conservative” 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 assessing 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 aversion 
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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 “protective”) 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 probability, 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 distributions were developed from available 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 available to characterize the input distribution and its 

uncertainties. In such cases, values that might be considered conservative 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 assessment 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 uncertainty. Distinguishing between them 

quantitatively is important for determining opportunities to reduce uncertainty in risk 

estimates. EPA’s input distributions are intended 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 conditions 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 present at the site.” (page 4-27) As a result, 

EPA has not separated uncertainty and variability 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 manner that would allow for a clear quantitative 

distinction between the contribution of variability 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 possible 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 waste onsite are not commensurate with the expected 

improvements in the CCW risk assessment results, EPA will not pursue such analysis in the revised risk 

assessment. 
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8.3 Calculation of Percentile Probabilities 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 probability 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 simply 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 considered. For the hazard quotients, the full 

distribution is necessary because hazard quotients are not probabilities. A clear 

explanation of the use of estimated distributions 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 stability of the Monte Carlo results. In other 

words, is the number of simulations sufficient, 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 function of how frequently WMUs with each liner 

type were selected in the Monte Carlo simulation. 
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EPA Response 

Figure 2 shows average risk versus the number of simulations for arsenic under selected CCW disposal 

scenarios, with the green and pink lines depicting landfill 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 cases, and the 10,000 simulations performed 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 would change in the future, thus affecting 

the potentially exposed population (e.g., higher 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 future 
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61

generations. Even with no adverse health impacts, 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 larger than the volume of groundwater that is 

actually contaminated because the pattern of groundwater contamination downstream of a 

source is nearly impossible to completely characterize, 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 analysis 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 water and the detrimental effects of loosing 

valuable water supplies. 

Dr. Kerry Rowe, Queens University 

In the absence of an actual survey of water 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 uncertainties in this analysis” (RA, pg. 4-37) 

however the statement that “EPA believes that the MSW well distance distribution used 

is protective for CCW landfills and surface impoundments” (RA, pg. 4-37) appears to be 

nothing more than that – a belief – with no justification 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 given 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 protection 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 that 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 consider potential impacts on the availability of 

groundwater in the future (e.g., contaminated groundwater becoming unsuitable for 

consumption), but the results do clearly indicate 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 quantitative 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 location, 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 distances to receptor wells used is similar 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. 
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9.0 COMMENTS RELATED TO THE DOCUMENTATION AND MISCELLANEOUS 

COMMENTS 

The peer review comments relating to documentation 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 Reproduce 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 are 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 provided. 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 insufficient 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 also 
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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 assessment 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 Carlo simulation as the source data derived 

from the EPRI survey, explaining that these data were combined with data from EPA’s 

CCW constituent database to conduct a simulation 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 the inputs EPA used are directly discernable from the 

appendices. Constituent concentrations and leachate values 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 MINTEQA2 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 sufficient for duplication of the results, and can provide 

the input and output files in Microsoft Access format upon request. 

9.2 Graphical Input/Output Distributions, 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 required for a PRA, including exposure input 

distributions, models used to predict surface water, fish, and groundwater concentrations, 

and models used to predict noncancer hazard quotients and cancer risks. EPA discusses 

many of the most important assumptions inherent 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-fit 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. 
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64

EPA explains that HQs for the child resident were consistently higher than HQs for the 

adult resident; therefore, only child HQs are presented. 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 documentation 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 online. The sensitivity analyses are an integral 

part of the PRA, vital to proper interpretation and use of results, and should not be 

reported in a separate document. 

EPA’s summary is not sufficient to understand the full scope of the sensitivity analysis 

effort, the procedures that EPA used, and extent to which EPA was able to distinguish 

influential sources of variability from influential sources of uncertainty. EPA excluded 

human exposure factors from the sensitivity analysis, describing them as sources of 

variability that could not be changed or managed 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 sensitivity 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 analysis of risk susceptibility patterns—i.e., what 

conditions lead to more or less relative risk. 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 where the risks are the highest, and then to map 
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65

where these locations are most likely to exist. This could be accomplished simply by 

taking the top, say 20%, of realizations with the worst risk and analyzing what these 

realizations have in common. Some characteristics 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 construction have a profound impact on the WMU 

infiltration rate for landfills (correctly identified 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 options for CCW (Objective 2), more cost 

effective management strategies may be developed 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 pathways, 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 include aquifer hydraulic conductivity and gradient, 

adsorption isotherm coefficient, depth to groundwater, receptor well distance, and waterbody flow rate. 

Although EPA did not analyze the top 20% of model runs as the commenter suggests, the sensitivity 

analysis results confirm that higher input values for these 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 climate variables as the commenter suggests. 

9.5 Discussion of Worst-Case Scenarios 

Comment Summary 

Dr. Rowe emphasized that worst-case scenarios may exceed the risks modeled in the risk 

assessment. 

Dr. Kerry Rowe, Queens University 

The discussion of receptor populations is considered 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 these circumstances, site specific studies 

would be warranted. These situations are unlikely to occur often but the possibility 

should be recognized in the development of any regulations based on the RA. For 

example, the consumption of fish assumes recreational fishing and this is, in most cases, 

likely to be reasonable. However if the receptor water body was used by subsistence fish 

consumers (native harvesters as identified in the comments by Dr. Foran) then the risks 

could be substantially higher than modeled. 
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66

EPA Response 

EPA acknowledges that a range of outcomes may occur. The draft risk assessment only showed the 

results for the 50th and 90th percentiles. Nevertheless, EPA appreciates the importance of site-specific 

analysis. Therefore, the risk assessment 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 the constituent-specific approach 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 water, as shown by exceedances of secondary MCLs or 

state standards. However, EPA modeled only the constituents most likely to result in human health or 

ecological risks, and because secondary MCLs are 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 background concentrations of CCW constituents in 

the absence of a release from WMUs. Such discussion would improve the risk 

assessment. 

EPA Response 

EPA acknowledges that a discussion of background concentrations 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. Although 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 the incremental increase in risk associated with 

constituents in wastes when conducting risk assessments under RCRA. 
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