2016-04-13 Amended_Cosler_Expert_Report_Allen

Amended Expert Report of
Douglas J. Cosler, Ph.D., P.E.
Chemical Hydrogeologist
Adaptive Groundwater Solutions LLC
Charlotte, North Carolina
Allen Steam Station Ash Basins
Belmont, North Carolina
April 13, 2016

Introduction
Site Backqround
The Allen Steam Station is a five -unit, coal-fired generating station owned by Duke Energy and located on
a 1,009-acre site near Belmont in Gaston County, North Carolina, adjacent to the west bank of the
Catawba River (Lake Wylie). Coal combustion residuals ("coal ash") have been disposed since 1957 in
an ash basin system located south of the station, consisting of an unlined active ash basin (formed in
1973) and an inactive ash basin. The ash basin system is generally located in historical depressions
formed from tributaries that flowed toward the Catawba River. For example, HDR notes in CSA Section
12.2.2.15 that the southern portion of the active ash basin was constructed over two streams. The
inactive ash basin consists of: two unlined dry ash storage areas (1996); two unlined structural fill units
(2003-2009); and a double -lined dry ash landfill (2009), commonly referred to as the Retired Ash Basin
(RAB) Ash Landfill. The RAB is located on top of the inactive ash basin immediately northeast of the
active ash basin. The ash storage areas and structural fill units overlie the western part of the inactive
ash basin.
Duke Energy performed biannual voluntary groundwater monitoring at the site from May 2004 to
November 2010 and NPDES permit -required monitoring starting in March 2011. Groundwater sampling
results at Allen indicate exceedances of 15A NCAC 02L.0202 Groundwater Quality Standards (2L
Standards). In response to this, the North Carolina Department of Environmental Quality (NC DEQ)
required Duke Energy to perform a groundwater assessment at the site and prepare a
Comprehensive Site Assessment (CSA) report. The Coal Ash Management Act of 2014 (CAMA) also
required owners of surface impoundments containing coal combustion residuals (CCR) to conduct
groundwater monitoring and assessment and prepare a CSA report. The August 2015 CSA prepared by
HDR Engineering, Inc. of the Carolinas (HDR) for the Allen site determined that ash handling and storage
at the Allen site have impacted soil and groundwater beneath and downgradient from the ash basin
system. The CSA report identified Constituents of Interest (COI) considered to be associated with
potential impacts to soil and groundwater from the ash basin and assessed COI concentration
distributions in soil, groundwater, and seeps.
CAMA also requires the submittal of a Corrective Action Plan (CAP); the CAP for the Allen site consists
of two parts. CAP Part 1 (submitted to DEQ in November 2015) provides a summary of CSA findings,
further evaluation and selection of COI, a site conceptual model (SCM), the development of groundwater
flow and chemical transport models of the site, presentation and analysis of the results of the modeling,
and a quantitative analysis of groundwater and surface water interactions. The CAP Part 2 contains
proposed remedial methods for achieving groundwater quality restoration, conceptual plans for
recommended corrective action, proposed future monitoring plans, and a risk assessment.
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Information Reviewed
My opinions are based upon an analysis and technical review of (i) hydrogeologic and chemical data
collected at the Allen site; (ii) the analyses, interpretations, and conclusions presented in site -related
technical documents and reports; (iii) the groundwater flow and chemical transport models constructed
for the site (including model development, calibration, and simulations of remedial alternatives); (iv) the
effectiveness of proposed remedial alternatives to achieve groundwater quality restoration; and (v)
proposed future site monitoring. This amended report contains additional opinions based on my review of
the recently -issued CAP Part 2 report. These opinions are subject to change as new information
becomes available.
As a basis for forming my opinions I reviewed the following documents and associated appendices:
(1) Comprehensive Site Assessment Report, Allen Steam Station Ash Basin (August 23, 2015);
(2) Corrective Action Plan, Part 1, Allen Steam Station Ash Basin (November 20, 2015);
(3) Corrective Action Plan, Part 2, Allen Steam Station Ash Basin (February 19, 2016);
(4) Miscellaneous historical groundwater and soil concentration data for the Allen site collected prior to
the CSA; and
(5) Specific references cited in and listed at the end of this report
Professional Qualifications
I have advanced graduate degrees in Hydrogeology (Ph.D. Degree from The Ohio State University) and
Civil and Environmental Engineering (Civil Engineer Degree from the Massachusetts Institute of
Technology), and M.S. and B.S. degrees from Ohio State in Civil and Environmental Engineering. I have
36 years of experience as a chemical hydrogeologist and environmental engineer investigating and
performing data analyses and computer modeling for a wide variety of projects. These projects include:
investigation, remediation, and regulation of Superfund, RCRA, and other hazardous waste sites involving
overburden and bedrock aquifers; ground water flow and chemical transport model development; natural
attenuation/biodegradation assessments for chlorinated solvent and petroleum contamination sites;
volatile organic compound vapor migration and exposure assessment; exposure modeling for health risk
assessments; hydrologic impact assessment for minerals and coal mining; leachate collection system
modeling and design for mine tailings disposal impoundments; and expert witness testimony and
litigation support. I also develop commercial groundwater flow and chemical transport modeling software
for the environmental industry.
The types of sites I have investigated include: landfills, mining operations, manufactured gas plants,
wood -treating facilities, chemical plants, water supply well fields, gasoline and fuel oil storage/delivery
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facilities, nuclear waste disposal sites, hazardous waste incinerators, and various industrial facilities. I
have investigated the following dissolved, nonaqueous-phase (LNAPL/DNAPL), and vapor -phase
contaminants: chlorinated solvents, various metals, gasoline and fuel oil constituents, wood -treating
products, coal tars, polychlorinated biphenyls, pesticides, dioxins and furans, phenolic compounds, flame
retardants (PBDE), phthalates, radionuclides, and biological constituents.
Summary of Opinions
The following is a brief summary of the opinions developed in my report:
• A total of 44 Compliance Boundary groundwater samples exceeded North Carolina groundwater
standards for these COI: boron, chromium, cobalt, iron, manganese, sulfate, total dissolved
solids, and vanadium. Of these 44 exceedances, 29 were greater than the proposed provisional
background concentrations by HDR, which exaggerate background levels;
• Most of the background wells are either likely to be downgradient from coal -ash source areas, or
appear to be downgradient from coal ash, and data from these wells overestimate actual site
background concentrations;
• The statistical analyses of background groundwater concentrations at the Allen site (well AB-1 R)
are invalid due to the characteristically slow rate of COI migration in groundwater;
• There is a significant risk of chemical migration from the ash basin to neighboring private and
public water supply wells in fractured bedrock;
• Major limitations of the CAP groundwater flow and chemical transport models prevent simulation
and analysis of off -site migration;
• The CAP Closure Scenario simulations greatly underestimate (by factors of 10 or more) the time
frames required to achieve meaningful groundwater concentration reductions in response to
remedial actions;
• For either the Existing Condition or Cap -in -Place Scenario groundwater concentrations of coal -
ash constituents much higher than background levels will continue to exceed North Carolina
groundwater standards at the Compliance Boundary because saturated coal -ash material and
secondary sources will remain in place;
• Source -area mass removal included in the Excavation Scenario results in COI concentration
reductions at the Compliance Boundary that are generally two to four times greater compared to
Cap -in -Place, best reduces impacts to surface water, and reduces cleanup times by a factor of
2.5-5. Additional excavation of secondary sources would further accelerate concentration
reductions;
• CSA data show multiple exceedances of groundwater standards in bedrock not only at the
compliance boundary but also inside the CB. However, the CAP Closure Scenarios do not
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address either concentration reduction or off -site chemical migration control in the fractured
bedrock aquifer;
• Due to an incorrect boundary -condition representation of the active ash basin, the CAP models
underestimate by more than a factor of two both the mass loading of COI into the Catawba River
and the corresponding Catawba River water concentrations (attributable to coal ash ponds)
estimated by the groundwater/surface-water mixing model;
• The CAP Part 2 geochemical modeling and monitored natural attenuation (MNA) evaluations do
not provide the required quantitative analyses of COI attenuation rates necessary to support MNA
as a viable corrective action. The CAP 2 chemical transport modeling, which included attenuation
by sorption, demonstrated that MNA is not an effective remedial option for several COI (e.g.,
antimony, boron, chromium, hexavalent chromium, cobalt, and vanadium);
• Future Compliance Monitoring at the Allen site should include much more closely -spaced
Compliance Wells to provide more accurate detection, and groundwater sampling frequency
should be re-evaluated to allow valid statistical analyses of concentration variations.
Hydrogeology of the Allen Site
Introduction
The groundwater system at the Allen site is an unconfined, connected system consisting of three basic
flow layers: shallow, deep, and fractured bedrock. The shallow and deep layers consist of residual soil,
saprolite (clay and coarser granular material formed by chemical weathering of bedrock), and weathered
fractured rock (regolith). A transition zone at the base of the regolith is also present and consists of
partially-weathered/fractured bedrock and lesser amounts of saprolite. The ash basins overlie native soil
and was constructed in historical drainage features formed from tributaries that flowed toward the
Catawba River using earthen embankment dams and dikes. HDR notes in CSA Section 12.2.2.15 that
the southern portion of the active ash basin was constructed over two streams. A large percentage of the
coal ash lies below the groundwater table and is saturated. Groundwater flow through saturated coal ash
and downward infiltration of rainwater through unsaturated coal ash leach COI into the subsurface
beneath the basin and via seeps through the embankments.
As described by HDR, groundwater flow in all three layers within the site boundary is generally from west
to east toward the Catawba River (Lake Wylie). Vertical groundwater flow between the three layers also
occurs, and surface water ponding in the active ash basin effects flow directions locally. Near the
western limits of the ash basin system is a groundwater "divide" where the flow direction changes from
west to east to a general westerly flow direction. The location of this divide and the nature of groundwater
flow west of the ash basin system were data gaps in the CSA and CAP Part 1 investigations; however,
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the CAP Part 2 did not further address this issue. On the eastern site boundary the CSA and CAP Parts
1 and 2 investigations assumed that all groundwater at the Allen site (overburden and bedrock aquifers)
discharges into the Catawba River. However, these studies did not collect hydrogeologic data or perform
data analyses or groundwater flow modeling to support this assumption. The CSA and CAP Parts 1 and
2 also did not analyze potential changes to site groundwater flow directions, or the risk of off -site
migration of COI in the overburden or bedrock aquifers, caused by groundwater extraction from numerous
private and public water supply wells located close to the site boundaries and near the eastern shore of
the river.
My report begins with a discussion of significant errors in CSA data analysis and conceptual model
development that contradict HDR's interpretation of three-dimensional groundwater flow patterns at the
Allen site. This is followed by a presentation and discussion of measured exceedances of North Carolina
groundwater standards at multiple locations on the ash basin compliance boundary. I then address
several limitations of the CAP Parts 1 and 2 groundwater flow and chemical transport models and identify
various model input data errors. Finally, I present my evaluations of the CAP Closure Scenario
simulations and provide my opinions regarding the effectiveness of various remedial alternatives for
restoring groundwater quality to North Carolina standards.
Errors in Hydraulic Conductivity Test Analyses
Throughout the CSA and CAP reports HDR provides interpretations and conclusions regarding the
horizontal and vertical variations of groundwater flow directions and rates, and the fate and transport of
COI dissolved in groundwater. The most important site -specific parameter that controls these time -
dependent flow and transport mechanisms is the hydraulic conductivity (also referred to as "permeability")
of the underlying soils and fractured bedrock (Bear, 1979). Hydraulic conductivity (length/time) is a
media -specific measure of the rate at which water can flow through a porous (soil) or fractured (bedrock)
porous medium. Groundwater flow and chemical transport rates are directly proportional to the product of
hydraulic conductivity and the hydraulic gradient (hydraulic head difference between two points divided by
the separation distance; e.g., the water table elevation slope at the Allen site). Therefore, accurate
measurement of hydraulic conductivity is critical for understanding the current and future distributions of
COI in soil and groundwater and for evaluating the effectiveness (e.g., cleanup times) of alternative
remedial measures.
In addition, the contrast in hydraulic conductivity between adjacent hydrogeologic units is the key factor in
determining three-dimensional groundwater flow directions and the ultimate fate of dissolved COI. For
example, at the Allen site accurate measurement of hydraulic conductivity is critical in evaluating the
potential for: downward chemical migration into the fractured bedrock unit, off -site COI migration in the
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overburden (soil) or fractured bedrock aquifers, groundwater flow and COI transport into or beneath the
Catawba River.
A slug test is one of the standard field methods for measuring hydraulic conductivity (K) using a soil
boring or installed monitoring well. Slug tests were performed in most of the overburden and bedrock
wells at the Allen site. In this test the static water level in the open hole (boring) or well casing is
suddenly increased or decreased and the resulting transient change in water level is recorded. Two
commonly -used techniques for quickly changing the water level are the introduction (increases the water
level) and removal (decreases the water level) of a solid rod, or "slug" into the boring or well casing.
These tests are called "falling -head" and "rising -head" tests, respectively. Higher rates of water -level
recovery correspond to higher values of K. The measurements of water level versus time are analyzed
using mathematical models of the groundwater flow hydraulics and information regarding the well
installation (e.g., length of the slotted monitoring well screen and well casing diameter) to compute an
estimate of K.
As discussed below, HDR made significant errors in all of their analyses of field slug test data. Their
analysis errors caused the reported (CSA report) slug test hydraulic conductivity values to be as great as
a factor of 50 smaller than the correct K values. I discuss the impacts of these major errors on HDR's
groundwater flow and chemical transport assessments and the CAP modeling throughout the remainder
of my report.
Bedrock Slug Tests
HDR made major mistakes in computing the bedrock hydraulic conductivity values using the field
monitoring -well slug -test results (CSA Appendix H). The Barker and Black (1983) slug test analysis
method used by HDR for bedrock wells was misapplied in all cases. The Barker -Black method is an
analytical solution that simulates water level changes in a well that fully penetrates a fractured isotropic
aquifer. The Barker -Black solution does not incorporate the well screen length as a parameter, as the
Bouwer-Rice (1976) method does (see following section). However, the screen length (length of the
slotted, or open, portion of the monitoring well) largely controls the hydraulic head changes in a partially -
penetrating (i.e., monitoring) well, along with the horizontal and vertical hydraulic conductivities of the
aquifer (Hantush, 1964; Bear, 1979). The ratio of the aquifer thickness to screen length is also
important.
As an illustration, in bedrock monitoring well GWA-3BR HDR used a well screen length of 248 feet
("saturated thickness" in their Appendix H Barker -Black analysis), but the true screen length in this well is
only five feet. Therefore, the Barker -Black model "thinks" a 248-foot thick aquifer is causing the water -
level change during the slug test, but the actual effective saturated thickness is about a factor of 50
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smaller (i.e., on the order of the five-foot screen length). Using five feet for the saturated thickness in the
GWA-3BR slug test analysis would yield a true K estimate that is about a factor of 50 higher (-- 248/5).
More accurate analyses of the bedrock slug tests also need to account for partial -penetration effects (i.e.,
the hydraulic effects of a pumping well with a "short" well screen located in a thicker aquifer).
Due to these errors the bedrock slug test results presented in the CSA are on the order of a factor of 50
too low. In the CAP flow model the bedrock is assumed to be factors of 40 to 4,000 lower than the
overburden aquifer in different areas (CAP Appendix C, Table 2). When these errors are corrected the
overall mean hydraulic conductivity for the bedrock unit is similar in magnitude to the overburden soils
and transition zone (CSA Table 11-10). If the mean bedrock permeability was actually 50-100 times
lower than the overburden aquifer there would be significant resistance to overlying groundwater and
dissolved COI migrating into the bedrock unit and flow and transport would tend to preferentially flow in
the overburden aquifer in a more predominantly horizontal direction (e.g., parallel to the bedrock surface).
However, that is not the case at the Allen site because the bedrock and overburden permeabilities are
similar in magnitude and downward groundwater flow in the upper portion of the bedrock unit is not
restricted. Moreover, the Allen site conceptual hydrogeologic model, the CSA field investigation and data
analyses, and the CAP flow and transport models were based on this incorrect bedrock hydraulic
conductivity assessment. I discuss these issues further in the "groundwater flow" section of my report.
Overburden Slug Tests
HDR analyzed all of the CSA overburden slug tests in shallow and deep wells with the Bouwer-Rice
(1976) method using a vertical anisotropy, Av = Knorizonrat / Kverticat , that is a factor of five lower than the
values presented in the CSA report (e.g., compare geometric mean values in CSA Tables 11-10 and 11-
11) and used in the CAP modeling (e.g., CAP 1 report Appendix C, Table 2). The values of Av in the CSA
and CAP modeling are generally on the order of ten for overburden soils, whereas the Bouwer-Rice slug
test analyses assumed Av = 2 (CSA Appendix H). Using the correct vertical anisotropy, Av = 10,
increases all of the measured overburden hydraulic conductivity values (CSA Table 11-4) by about 25
percent, depending on how the slug -test radius of influence was computed.
Since every reported overburden K value in the CSA report (at least for new shallow and deep wells) is
approximately 25 percent too low, the actual average chemical transport rates in overburden soils are
about 25 percent greater than reported. This site -wide data reduction error also affects the CAP 1 and 2
flow and transport model calibrations. For example, HDR significantly reduced laboratory measurements
of the soil -water partition coefficient, Kd, for various COI during the transport model calibration based on
comparisons of observed and simulated chemical migration rates. However, if the correct (i.e., higher)
overburden K values had been used in the model calibration the Kd values would not have been reduced
as much (compared to laboratory values). The reason for this is, assuming linear equilibrium partitioning
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of COI with soil, the chemical migration rate is proportional to KI Kd (except for Kd << 1). The CAP 1
transport model history matching indicated that the simulated transport rate was too low, so the model
developers reduced the model Kd. In other words, the reductions in calibrated Kd values would not have
been as great if the correct (higher) K values were used in the first place. As discussed below, the CAP
Part 2 transport modeling used Kd values that are generally a factor of about 10 larger than the CAP 1
values; however, the CAP 2 Kd's are still on the order of 10 times smaller than the measured site -
specific Kd's reported in CAP 1 Appendix D and CAP 2 Appendix C. COI sorption to soil is important
because, as discussed below, aquifer cleanup times (i.e., chemical flushing rates) are generally
proportional to the chemical retardation factor, which is directly proportional to Kd, except when Kd << 1
(Zheng et al., 1991).
Groundwater Flow
Throughout the CSA and CAP reports HDR made several critical assumptions, not supported by data,
regarding the horizontal and vertical groundwater flow directions near the boundaries of the Allen site
which impacted their conclusions regarding the ultimate discharge locations for site groundwater and
dissolved COI. Two examples discussed in this section are (i) the relationship between site groundwater
and the Catawba River and (ii) groundwater flow directions and the potential for offsite migration of COI.
Catawba River and the LeGrand Conceptual Model
All groundwater at the Allen site was apparently assumed to discharge into the Catawba River according
to a generalized conceptual model (LeGrand, 2004) before actual site -specific hydrogeologic data were
analyzed. Statements to this effect were made at numerous points in the CSA and CAP reports.
However, HDR did not present any site -specific data analyses or groundwater flow modeling that would
support this assumption in either report. In fact, as discussed below, the boundary conditions for the CAP
Parts 1 and 2 flow models effectively forced site groundwater to discharge into the river at the
downgradient model boundary.
HDR continued to state this assumption in the CAP 2 report (e.g., Section 3.3.2) even though the
consistently downward groundwater flow components next to the Catawba River show that this
assumption is incorrect (refer to the deep/bedrock aquifer vertical hydraulic gradients measured during
the CSA and CAP Part 2 investigations). In CAP 2 Section 3.3.2, HDR also states that "The Catawba
River serves as the lower hydrologic boundary for groundwater within the shallow, deep, and bedrock
flow layers at the site." However, the river cannot be the "lower hydrologic boundary" for the deep and
bedrock layers when the measured vertical flow direction in these layers is consistently downward, which
demonstrates that HDR has not delineated this inferred "lower boundary" used in the CAP models. HDR
further states in CAP 2 Section 3.3.2 that "the approximate vertical extent of the groundwater impacts is
generally limited to the shallow and deep flow layers." This statement ignores that fact that groundwater
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flow across the site is consistently downward from the impacted deep flow layer to the highly -fractured
bedrock aquifer and that, as discussed below, 12 exceedances of North Carolina 2L and/or IMAC
groundwater standards (and greater than background concentrations) were measured in samples
collected from bedrock wells located inside the Compliance Boundary.
The LeGrand (2004) guidance document presents a general discussion of groundwater flow patterns that
may occur near streams in the Piedmont and Mountain Region of North Carolina based on ground
surface elevations (i.e., site topography and surface watershed boundaries). However, surface water and
groundwater watersheds commonly do not coincide (Winter et al., 2003). Further, groundwater flow
patterns and rates in bedrock have been found to be poorly related to topographic characteristics (Yin
and Brook, 1992). LeGrand does not present or derive any mathematical equations or quantitative
relationships for groundwater flow near rivers or streams. The author emphasizes that site -specific data
must be collected in order to correctly evaluate river inflow or outflow. In strong contrast to the LeGrand
generalizations, numerous detailed and sophisticated mathematical (analytical and numerical) river -
aquifer models and highly -monitored field studies have been published in the scientific and engineering
literature in the past several decades. What these investigations and applied hydraulic models show is
that the water flow rate into or out of a river or stream and the depth of hydraulic influence within an
underlying aquifer are highly sensitive to several factors, including: the transient river water surface
elevation and slope; river bed topography; bed permeability and thickness; horizontal and vertical
permeability (and thickness) of the different hydrogeologic units underlying the river; transient horizontal
and vertical hydraulic head variations in groundwater beneath and near the river; and groundwater
extraction rates and screen elevations for neighboring pumping wells (e.g., Simon et al. 2015; McDonald
and Harbaugh, 1988; Bear, 1979; Hantush, 1964).
The CSA investigation did not: measure river bed permeability or thickness; characterize the river
bathymetry; monitor transient water surface elevation variations at more than one location (one average
value was used); collect river bed hydraulic gradient data; measure horizontal or vertical overburden or
bedrock permeability beneath or on the east side of the river; characterize the geology beneath or east of
the river; measure hydraulic heads in the overburden or bedrock beneath or east of the river; or consider
the hydraulic effects of groundwater extraction from nearby water supply wells (e.g., the numerous private
homes located on the east bank of the Catawba River, near the southeast corner of the site, and west of
the property boundary). Much of the data that were collected in the CSA contradict the LeGrand
hypothesis at the Allen site. For example, strong downward flow components from deep to bedrock wells
were measured at each deep/bedrock monitoring cluster (GWA-5, GWA-3, and GWA-1) next to the
Catawba River (compare CSA Figures 6-6 and 6-7), and the following shallow -deep well clusters on the
west river bank: GWA-4, AB-9, and AB-10 (CSA Figures 6-5/6-6). Downward groundwater flow was also
measured at several other locations across the site (CSA Table 11-14). Water -level measurements
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collected in September 2015 for the CAP Part 2 investigation [CAP 2 Figures 2-2 (shallow), 2-3 (deep),
and 2-4 (bedrock)] demonstrated downward groundwater flow from the deep to bedrock aquifers at all
locations across the Allen site with the exception of monitoring well clusters AB-25D/BR and AB-35D/BR.
Maps of vertical hydraulic gradient variations (e.g., contour maps) were not generated for the CSA or
CAP Part 2, and HDR did not discuss the significance of downward hydraulic gradients next to the river
and at most other deep/bedrock monitoring well clusters. These downward groundwater flow
measurements are consistent with the hydraulic conductivities of the bedrock and overburden being of
similar magnitude, as discussed above.
The strong and consistent measured downward groundwater flow components right next to the Catawba
River and at most other well clusters indicate that site groundwater is entering the deep fractured bedrock
unit at the downgradient property boundary (and many other areas of the site) and that not all of the site
groundwater discharges into the river as the site Conceptual Model and the CAP flow and transport
models assume. The downward flow into bedrock at the downgradient property boundary may also be
due in part to groundwater extraction from private bedrock water supply wells located along the eastern
shore of the river, but in the CSA and CAP investigations HDR assumed these factors related to the
potential for off -site COI migration beneath the river were not important and did not evaluate them. The
assumption that all site groundwater discharges into the Catawba River may also be related to the sparse
number of bedrock monitoring wells (two) that were installed along the entire downgradient property
boundary.
Groundwater Flow Directions
The CSA assumptions and analysis errors discussed above have had a strong effect on: the Conceptual
Model development; the site hydrogeologic and COI transport assessment; the construction/calibration
of the CAP flow and transport models; and the simulations of CAP Close Scenarios. The hydrogeologic
assumptions should have been carefully evaluated and tested during the performance of the CSA and as
part of the CAP groundwater flow model construction and calibration to determine whether they were
valid. Instead, the hypotheses appear to have effectively guided the model development and led to
inaccurate interpretations.
As an illustration, the slug -test analysis mistakes have resulted in several inaccurate hydrogeologic
interpretations and conclusions in the Allen CSA with regard to groundwater flow and chemical transport
rates and directions (horizontal and vertical). Because the permeability of the weathered bedrock is
similar to the overlying soils the CSA and CAP interpretations that the bedrock acts as a lower confining
layer for groundwater flow and chemical transport is incorrect. As further discussed below in the model
review section, the much higher (after correction for errors) bedrock permeability also generally increases
the potential for off -site COI migration toward private and public water supply wells. Therefore, the CSA
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and CAP conclusions that (i) all site groundwater discharges into the Catawba River and (ii) groundwater
and dissolved coal -ash constituents are restricted from migrating to the west toward the residential and
water supply wells are not consistent with the data.
Exceedances of Groundwater Standards
In this section I compare measured groundwater concentrations in shallow, deep, and bedrock
groundwater samples to North Carolina 2L and IMAC standards and show the following: (i) 34 measured
exceedances for several COI at multiple locations on the Compliance Boundary (CB); (ii) an additional 10
exceedances at CB locations beneath the Catawba River based on chemical transport modeling I
performed; (iii) 29 of the 44 Compliance Boundary exceedances were greater than the proposed
provisional background concentrations (PPBC) by HDR; (iv) 20 of the 44 Compliance Boundary
exceedances were greater than the maximum background concentration in the same hydrogeologic unit
(e.g., shallow, deep, or bedrock); (v) 13 additional exceedances were measured in wells located on the
Duke Energy property boundary (PB) in areas where the CB and PB do not coincide, as drawn by HDR;
(vi) 12 additional exceedances were observed in wells screened in the highly -permeable fractured
bedrock unit underlying the ash basin and located inside the CB; (vii) several background wells are likely
downgradient from coal ash and cannot be relied upon to provide accurate COI background levels; (viii)
excluding these questionable background wells which I believe to be downgradient from coal ash, 32 of
the 44 Compliance Boundary exceedances were greater than the maximum background concentration in
the same hydrogeologic unit; and (ix) the statistical analyses of groundwater concentrations at well AB-
1 R for purposes of defining background levels were performed incorrectly [HDR also believes AB-1 R may
be downgradient from COI discharging from the inactive ash basin (CSA Report Section 10.2); in the
CAP 2 report HDR states that AB-1 R will no longer be used as a background well due to increasing COI
concentration trends at this location].
Throughout this report I reference the ash basin compliance boundary and the Duke Energy property
boundary for the Allen site as drawn on maps developed by HDR (e.g., CSA Figure 6-2). My reference to
the "compliance boundary" is only for identification purposes and not an opinion that this boundary as
drawn by HDR is accurate or legally correct.
Summary of Exceedances
Table 1 summarizes exceedances of 2L or IMAC standards in shallow, deep, and bedrock groundwater
samples obtained from monitoring wells located: (i) on the Ash Basin Compliance Boundary (CB) as
drawn by HDR; (ii) on the Duke Energy property boundary (PB), where the CB and PB do not coincide;
(iii) bedrock wells (BR) located inside the CB; and (iv) modeled Compliance Boundary concentrations
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(CBM), using modeling techniques described below. The proposed provisional background
concentrations (PPBC) by HDR are also listed in Table 1.
A total of 34 Compliance Boundary groundwater samples exceeded North Carolina groundwater
standards for these COI: boron, chromium, cobalt, iron, manganese, sulfate, total dissolved solids, and
vanadium. I estimated an additional 10 exceedances at CB locations beneath the Catawba River based
on chemical transport modeling and measured upgradient concentrations (CBM). In addition, 12
exceedances were observed in wells screened in the highly fractured bedrock unit located inside the CB.
Thirteen more exceedances were measured in wells located on the Property Boundary in areas where
the CB and PB do not coincide, according to maps provided with the CSA and CAP. Note that the CAP 2
transport model also predicts year 2015 Compliance Boundary exceedances for these COI: antimony
[shallow(S), deep(D), bedrock(BR)]; boron (S,D); chromium (S,D,BR); hexavalent chromium (S,D,BR);
cobalt (S,D,BR); sulfate (S,D,BR); and vanadium (S,D,BR) (CAP 2 report, Table 4-1).
Excluding questionable background wells which may be downgradient from coal ash, (footnote "g" in
Table 1; refer to discussion below), 32 of the 44 (measured plus modeled) Compliance Boundary
exceedances were greater than background levels (from the same hydrogeologic unit) for a particular
constituent. Nine of the 13 PB exceedances were greater than background levels. Because the bedrock
background well BG-2BR may be downgradient from coal ash, it appears that all of the 12 bedrock
exceedances were greater than BR background levels.
Ignoring the potential problems with background well (BG) locations (i.e., using all background wells), 20
of the 44 Compliance Boundary exceedances were greater than any BG concentration (from the same
hydrogeologic unit) for a particular constituent. Five of the 13 PB exceedances were greater than
maximum background levels. The concentrations for 10 of the 12 bedrock exceedances were greater
than any BR background level.
A total of 29 of the 44 Compliance Boundary exceedances were greater than the proposed provisional
background concentrations (PPBC) by HDR, which exaggerate background levels due to their reliance on
data from monitoring wells that I believe are downgradient from coal ash. Seven of the 13 property
boundary exceedances were greater than the PPBC.
Note that the iso-concentration contours in all of the CSA Section 10 figures and Figure ES-3 are not
consistent, and are in many cases misleading, with regard to chemical transport mechanisms in the
subsurface. For example, the iso-concentration contours in Section 10 generally closely encircle a
monitoring well and infer no subsequent transport downgradient from the well location. This contouring
problem is especially prevalent near the downgradient Duke Energy property boundary which coincides
13

with the western shore of the Catawba River. Figure ES-3 is a good example of this practice. These
closed contours at the downgradient property boundary suggest that COI transport beyond the farthest
downgradient line of monitoring wells does not occur and that no COI reach the Compliance Boundary
located beneath the river (where there are no monitoring wells). This is not the case, of course, for
several COI as demonstrated in the following section, where COI exceedances at the Catawba River
section of the Compliance Boundary are demonstrated by modeling, and illustrated by most of the CAP
model simulated "existing conditions" plume maps in CSA Appendix C which contain 'open contours" at
the river shoreline. If the CAP model grid did not stop short of the Compliance Boundary (CB) the
Appendix C existing condition COI plumes would reach the CB in most cases.
Modeled Compliance Boundary Exceedances
I computed Compliance Boundary (CB) concentrations labeled "CBM" with footnote "e" in Table 1 using a
calibrated one-dimensional, analytical chemical transport model (van Genuchten and Alves, 1982;
Equation C5) because the CB at these locations was up to 500 feet offshore beneath the Catawba River.
The CAP transport model grid did not extend to this downgradient portion of the CB, and HDR did not
evaluate COI concentrations at this location. I calibrated the analytical model to chemical -specific site
conditions (i.e., determined model input parameter values) using CAP transport model simulated
concentration versus time curves for "Existing Conditions" (CAP report Appendix C). The analytical
model input parameters in my model were: groundwater pore velocity, chemical retardation factor, and
longitudinal dispersivity.
Figures 1 a and 1 b show the close agreement between the analytical model simulation results (solid lines)
for Boron (Well GWA-4S) and Cobalt (Well GWA-5S), respectively, and the CAP model predictions from
Appendix C (solid circles). For each constituent, I used the calibrated analytical model to compute the
concentration versus time curve immediately downgradient at the Compliance Boundary (dashed lines).
The modeled CBM concentrations in Table 1 are equal to the product of the measured monitoring well
concentration and the ratio of the simulated CB and simulated monitoring well concentrations for 2015
(Figure 1).
Exceedances of Groundwater and Surface Water Standards in Seep Samples
Concentrations in several seep water samples exceeded relevant NCAC 2B, 2L and/or IMAC standards
for various COI (e.g., CSA Table 7-8, CSA Appendix F, CSA Figure 2-2, CAP Table 2-5.1 and 2-5.2). As
discussed in the introduction, the active ash basin at the Allen site is constructed above two historical
streams that are tributaries of the Catawba River (e.g., CSA Figure 2-3). Seeps S-3 and S-4, located
downgradient of the ash basin between the toe of dams and the Catawba River, coincide with the
downgradient portions of the historical streams visible on topographic maps (CAP Figure 2-2). Recent
surveys identified streams at these locations and at locations coinciding with seep locations S-1 and S-2,
14

as well as a wetland complex (CAP Figure 1-5). HDR describes seep locations S-1 through S-4 as
tributaries flowing toward the river (Lake Wylie) with well-defined stream channels of varying widths (2014
Seep Water and Monitoring Plan; Table 1).
Referring to my Table 1, 14 of the seep samples exceeded North Carolina groundwater standards (CSA
Table 7-8) for these COI: boron, iron, manganese, and vanadium. These samples are from seep
locations S-3, S-4, ANWW-002, ANWW-004, ANSW-001, and ANSW-015. The CAP report describes
seep locations S-1 and S-2 as "dry" and does not report sampling results, although other reports describe
streams with continuous flow at these locations. North Carolina surface water (213) standards have been
exceeded in seeps samples for these constituents (locations not specified): aluminum, copper, lead,
manganese, mercury, and total dissolved solids (CAP Table 2-5.2). CSA Appendix F notes that previous
seep sampling at locations S-1 through S-9 detected 2B exceedances for iron, manganese, zinc, and
thallium.
Determination of Background Concentrations
Background Well Locations
Most of the background wells at the Allen site have not been installed in locations that are clearly
upgradient from site -related coal -ash materials. For example, designated background well AB-1 R is likely
downgradient from the Inactive Ash Basin and Ash Storage Areas based on the hydraulic head
measurements at wells AB-38S and AB-1 R (e.g., the 622-, 623-, and 624-foot contours). In this regard,
Section 10.2 of the CSA report discusses the fact that a recent trend of increasing coal -ash constituent
concentrations has been observed at this location and acknowledges that AB-1 R may be downgradient
from the Inactive Ash Basin. The CSA report further recommends that data from this well needs to be
carefully evaluated to determine whether it can continue to be used for background monitoring. In the
CAP 2 report HDR states that AB-1 R will no longer be used as a background well due to these problems.
In addition, well BG-1 S is directly downgradient from the western half of the Active Ash Basin based on
the 636- and 637-foot hydraulic head contours in CSA Figure 6-5 (HDR did not draw these specific
contours). The light -blue "approximate groundwater flow direction" arrows in this area are incorrectly
drawn. Well BG-2S may be downgradient from the Inactive Ash Basin and the Ash Storage Areas
according to hydraulic head measurements in these areas.
Deep background well BG-1 D is likely downgradient from the western half of the Active Ash Basin based
on hydraulic head measurements in this area [e.g., the 636-foot hydraulic head contour in CSA Figure 6-6
(not drawn by HDR)]. The light -blue "approximate groundwater flow direction" arrows in this area are
15

incorrectly drawn. Well BG-21D may be downgradient from the Inactive Ash Basin and Ash Storage
Areas.
Bedrock background well BG-2BR may be downgradient from site -related coal -ash materials based on
the bedrock hydraulic head map (CSA Figure 6-7) and due to groundwater extraction from private and
public water supply wells located immediately west of the Allen site. As discussed throughout my report,
the CSA and CAP modeling failed to analyze the effects of off -site groundwater extraction.
Statistical Analyses of Background Concentrations
Appendix G of the CSA report presents statistical analyses of historical concentrations from shallow
Monitoring Well AB-1 R, which HDR described as following methods specified by the U.S. Environmental
Protection Agency (EPA, 2009), in an attempt to establish background groundwater concentrations for the
Allen site. As outlined in Sections 3.2.1 and 5.5.2 of the EPA guidance document these data must be
checked to ensure that they are statistically independent and exhibit no pairwise correlation.
Groundwater sampling data can be non -independent (i.e., autocorrelated) if the sampling frequency is too
high (i.e., time interval between sampling events is too small) compared to the chemical migration rate in
the aquifer (groundwater pore velocity divided by chemical retardation factor). Section 14 of the EPA
guidance presents methods for ensuring that the AB-1 R background data are not autocorrelated, but the
analyses in CSA Appendix G did not include evaluations for statistical independence.
As an illustration, "slow -moving" groundwater combined with high chemical retardation (i.e., large soil -
water partition coefficients, Kd), which is the case at the Allen site, can lead to the same general volume
of the chemical plume being repeatedly sampled when the monitoring events are closely spaced. For
example, the groundwater pore velocity (Vp) near well AB-1 R is on the order of 8 feet per year based on
the measured hydraulic conductivity (slug test results in CSA Table 11-5) and the horizontal hydraulic
gradients in this area (CSA Figure 6-5). Note that shallow pore velocities are generally a factor of 200
greater in many areas downgradient of the ash basin system due to much greater hydraulic gradients (-
20x larger) and larger hydraulic conductivity (— 10x greater) in this area. In addition, groundwater pore
velocities in deep overburden and in fractured bedrock are generally more than a factor of 1,000 greater
than velocities in the shallow overburden (CSA Table 11-13).
The retardation factors, Rd, based on laboratory Kd measurements (Kd — 10 cm3/g, or greater) are on the
order of 100 (or greater) for many of the COI (except conservative parameters such as sulfate and
boron). Accordingly, based on measured sorption values and average permeability values for the aquifer
the mean chemical migration rate (Vp /Rd) near background well AB-1 R is on the order of 0.1 ft/yr for
several of the non -conservative COI, assuming linear equilibrium sorption (refer to discussion below). For
the AB-1 R quarterly sampling, the non -conservative chemical migration distance between AB-1 R
16

sampling rounds (3 months) is less than 0.1 feet for some COI, which is approximately equal to the
sandpack diameter for the monitoring wells. Therefore, the AB-1 R background samples are basically
representative of the same volume of the plume (i.e., the vicinity of the sandpack, depending on the well
purge volume) for many COI, and the measured sample concentration changes analyzed in the CSA are
not due to actual chemical transport effects in the overburden aquifer. This means that the groundwater
samples are non -independent and that the statistical analyses of background concentrations at AB-1 R do
not satisfy the key requirements of the analysis method.
CAP Groundwater Flow Model Underestimates Potential for Off -Site Chemical Migration
My discussions in this section focus on limitations of the CAP Parts 1 and 2 groundwater flow models. I
focus specifically on model boundary conditions representing the Catawba River; the overall size of the
model grid and no -flow boundary conditions on the western, southern, and northern grid boundaries; the
misrepresentation of groundwater flow in the fractured bedrock aquifer; and the potential for off -site
groundwater flow in relation to groundwater extraction from numerous private and public water supply
wells located close to the model boundaries, but not incorporated into the flow model.
Catawba River Boundary Condition
The CAP Parts 1 and 2 groundwater flow models force all groundwater at the downgradient property
boundary to discharge directly into the Catawba River and significantly underestimate the potential for off -
site flow and chemical migration in fractured bedrock. [The Electric Power Research Institute (EPRI)
model review letter (October 20, 2015) refers to the use of drain (i.e., "leaky type") boundary conditions in
the flow model, through which 17 percent of the site groundwater discharges from the model, but the CAP
reports do not show their location or clearly discuss their hydraulic effect on the flow regime.] No -flow
boundary conditions defined along the entire western, northern, and southern model boundaries and the
bottom two layers (i.e., bedrock) of the eastern boundary prevent any off -site flow and chemical transport
in these areas (refer to Figures 1 and 4 in Appendix C of the CAP 1 Report). The bottom surface
(bedrock) of the flow model is also assumed to be a no -flow boundary even though the hydraulic
conductivity data and measured downward hydraulic gradients at most deep/bedrock monitoring well
clusters contradict this assumption. The only locations where groundwater and dissolved constituents are
allowed to leave the CAP models (other than the unexplained drain leakage at the top of the model) are
the vertical array of cells located above the bedrock unit along the eastern grid boundary, which are
assumed to represent the Catawba River; these cells are specified as constant -head boundary
conditions in which the head is uniform with depth.
17

This hydraulic representation of the Catawba River in the flow models is inaccurate for many reasons.
First, the river bottom is assumed to extend all the way through the overburden aquifer to the bedrock
surface, which is not the case. Second, groundwater flow beneath and adjacent to the river is assumed
to be horizontal with zero vertical flow component. Because this boundary condition does not allow
groundwater to flow vertically in areas beneath and near the river, the CAP models do not represent
actual site hydrologic conditions. Groundwater flow at the Allen site is not strictly horizontal and, as
discussed above, most of the deep/bedrock vertical hydraulic gradient measurements (including next to
the river) are downward. Third, as represented in the CAP models, neither the lower -permeability river
bed sediments nor the smaller vertically hydraulic conductivity of underlying soils restricts the potential
flow rate into or out of the river (i.e., in the CAP models a perfect hydraulic connection exists between the
aquifer and the Catawba River). The actual degree of aquifer -river hydraulic connection was not
evaluated in the CSA or CAP Parts 1/2. In summary, due to all of these factors the potential for site
groundwater and dissolved constituents to migrate off -site eastward beyond the Catawba River or
southward as underflow beneath the river cannot be evaluated with the CAP models. As I discuss in this
report, the consistent measured downward groundwater flow components from deep overburden into the
highly -permeable bedrock aquifer next to the Catawba River and at most other locations across the site
indicate that COI may be migrating off -site beneath the river, which is a major contradiction of the Allen
site conceptual model.
The CAP models should have represented the Catawba River using a "leaky -type" (i.e., river) boundary
condition in the top model layer (McDonald and Harbaugh, 1988), and the model grids should have
extended farther east so that HDR could have evaluated the above factors during model calibration and
sensitivity analyses. In their reviews of both the CAP 1 and 2 models (submitted with the CAP modeling
appendices), the Electric Power Research Institute third -party peer review team also concluded that the
Catawba River should be modeled as a leaky boundary condition instead of using constant heads. The
models also should have included groundwater extraction from the private water supply wells installed at
many points close to the eastern river bank. A river boundary condition incorporates the bed permeability
and thickness, the river water surface elevation, and the simulated hydraulic head in the aquifer (at the
base of the river bed) to dynamically specify a flux (flow rate per unit bed area) into or out of the
groundwater model depending on the head difference between the river and aquifer. Typically,
permeability and vertical hydraulic gradient measurements for the river bed (not collected in the CSA) and
flow model calibration (three-dimensional matching of simulated and measured hydraulic head
measurements in the aquifer) are used to determine a best -fit estimate of river bed conductance
(permeability divided by thickness) in the model. HDR did not perform this routine analysis.
18

Limitations of No -Flow Boundary Conditions and Small Model Domain Size
The limited areal extent and depth of the CAP Parts 1 and 2 flow and transport model grids prevent the
use of the models as unbiased computational tools that can be used to evaluate off -site migration of coal -
ash constituents. For example, the model grids should have extended farther west to include the South
Fork Catawba River and also incorporated groundwater extraction from off -site private -home and public
water -supply wells. The western no -flow boundary in the current CAP 1 and 2 models artificially prevents
any off -site flow or transport to the west in either the bedrock or overburden aquifers. The same is true
for the entire northern and southern model boundaries despite the fact that several private homes are
also located south of the Active Ash Basin and shallow, deep, and bedrock hydraulic head maps exhibit
southeasterly flow components in this area. Artificial limitations created by the eastern Catawba River
boundary condition are outlined above. In addition, the downgradient boundaries of the CAP 1 and 2 flow
and transport models do not extend to the Compliance Boundary (except for a small interval near Well
GWA-3), which prevents their use for estimating Compliance Boundary concentrations for various
remedial alternatives.
The bottom boundaries of the CAP models should extend much deeper because the hydraulic
conductivity of the fractured bedrock zone is of the same order of magnitude as the overburden soils
based on my corrections for slug test analysis errors. In the present configuration the lower boundaries of
the CAP Parts 1 and 2 model grids are only about 50 feet below the bedrock surface (Figure 3 in both the
CAP 1 & 2 modeling appendices). Because several bedrock wells were screened to this depth the
bedrock hydraulic conductivity data collected for the CSA demonstrate that imposing an impermeable
model boundary at this depth is incorrect. As discussed above, the strong measured downward hydraulic
gradients between deep and bedrock wells in both the June and September 2015 sampling rounds (e.g.,
Wells GWA-5, GWA-3, GWA-1 next to the Catawba River, and Well BG-2) demonstrate that vertical and
horizontal groundwater flow in bedrock is important, and these transport mechanisms need to be
accurately simulated in the CAP models in order to accurately assess the potential for off -site chemical
migration.
Model Significantly Underestimates Leakage Rate from Active Ash Basin
The CAP Part 2 flow model underestimates leachate discharge from the active ash basin by as much as
a factor of 100 in areas of ponded surface water (e.g., refer to CSA Figure 4-3). The CAP 1 model
underestimates active basin leakage by as much as a factor of 200. As shown in Figure 5 of CAP 1,
Appendix C, the CAP 1 flow models assumes a constant groundwater recharge rate (i.e., basin leakage
rate) equal to 8.5 inches/year in the active ash basin. In the CAP 2 flow model the active basin leakage
rate is assumed to be 14.5 inches/year (Figure 5 of CAP 2, Appendix B). However, CSA Figure 8-8.2
(cross-section F-F) shows that the vertical hydraulic gradient through the coal ash in the downgradient
portion of the active ash basin is on the order of unity. Using Darcy's law and the mean vertical coal -ash
19

permeability of 1.6E-4 cm/sec in CSA Table 11-11, the approximate vertical leakage rate out of the active
basin is about 2,000 inches/year near the Catawba River (i.e., — 140 times greater than the specified
recharge rate of 14.5 inches/year).
The CAP flow models should have represented ponded areas of the active ash basin as either constant -
head or leaky -type boundary conditions, which would have allowed the model to simulate a realistic
leakage rate for the active ash basin. The major discrepancies between the measured shallow hydraulic
head maps (CSA Figure 6-5 and CAP 2 Figure 2-2) and the CAP 1 and 2 simulated shallow head maps
(Figure 14 in CAP 1, Appendix C; Figure 14 in CAP 2, Appendix B) clearly show that the CAP flow
models significantly underestimate the hydraulic head beneath the active ash basin due to the fact that
the modeled leakage rate from the active basin is much too low.
Three related impacts of this incorrect active basin boundary condition are that the CAP models
significantly underestimate: (i) vertical groundwater flow rates (by on the order of a factor of 100) through
coal -ash source material in the vicinity of the downgradient portion of the active ash basin; (ii) horizontal
groundwater flow and chemical transport rates (— factor of two) downgradient from the active ash basin
(compare measured and simulated horizontal hydraulic gradients between the ash basin and the
Catawba River); and (iii) vertical flow rates from the overburden aquifer into the fractured bedrock unit
beneath ponded areas. This incorrect boundary condition representation of the active ash basin also
causes the CAP models to significantly underestimate (by more than a factor of two) both the mass
loading of COI into the Catawba River and the corresponding Catawba River surface water
concentrations (attributable to coal ash ponds) that HDR estimated with their mixing model (e.g., CAP 2
report Table 4-2 and Appendix D).
Off -Site Groundwater Extraction Ignored
The CSA and CAP Parts 1 and 2 failed to examine the strong potential for coal -ash constituents from the
Allen site to migrate with groundwater to private and public bedrock water supply wells located
immediately west, south, and east (near the eastern bank of the Catawba River) of the Duke Energy
property boundary. CSA Figure 4-2 shows the locations of private and public water supply wells near the
site. The basis of my opinion includes the following: hydraulic conductivity measurements for the
overburden and bedrock formations, including my corrections to the bedrock slug test analyses; three-
dimensional variations in measured hydraulic head in the bedrock and overburden units; groundwater
concentration data; and calculations of potential hydraulic head reductions (i.e., drawdown) that could be
caused by off -site groundwater extraction. As discussed throughout my report, neither the CSA nor CAP
Parts 1 or 2 investigations addressed the potential for off -site migration.
20

COI were detected in several water supply well samples (CSA Appendix B), but the CSA report did not
plot these detections on a map and did not discuss their possible relationship to the Allen site. Appendix
B also did not present the well construction details (e.g., well diameter and elevation range of the well
screen or open bedrock interval) so that well dilution effects and potential chemical transport pathways in
the bedrock unit could be evaluated. In addition, the CSA investigations and CAP Part 1 modeling did not
include these areas west of the Allen site. The CAP Part 2 flow model did include a small number (44 of
the 219 neighboring private wells and none of the four public water supply wells) of residential wells
located inside the undersized model domain (between Southpoint Road and the active ash basin), but the
CAP 2 modeling report (CAP 2, Appendix B) did not show simulated hydraulic head maps with these
residential wells pumping and did not provide any discussion or analyses of the potential for these wells
to capture COI dissolved in groundwater. The CAP Part 2 also did not increase the model grid size to
incorporate the large number of residential and water supply wells located immediately west, south, and
east of the site (eastern shoreline of the Catawba River); fix the boundary condition problems; or correct
the model input data errors so that the flow and transport models could be used to more accurately
analyze the potential for off -site chemical transport.
One of the major input data errors is the bedrock hydraulic conductivity, which is assumed in the CAP
Parts 1 and 2 flow models to be factors of 100 to 10,000 lower than the overburden aquifer in most areas
(CAP 1 Appendix C, Table 2; CAP 2 Appendix B, Table 2). A corrected analysis of the bedrock slug test
data shows that the mean bedrock permeability is approximately the same as the overburden
permeability. Therefore, downward groundwater flow from the overburden aquifer into the upper portion
of the bedrock unit is not restricted, as the CAP flow models represent, and the potential for off -site
chemical migration is underestimated by the models.
CAP Chemical Transport Modeling
Due to model calibration, model construction, and boundary -condition and input -data errors the CAP
models significantly underestimate remediation time frames. As discussed in this section, reasons for this
include underestimation of the chemical mass sorbed to soil, failure to account for slow chemical
desorption rates, inaccurate analyses of water -table lowering due to capping, and flaws in the transport
model calibration.
Soil -Water Partition Coefficients and Model Calibration
The fraction of chemical mass sorbed to soil can be represented by the soil -water partition coefficient, Kd
(Lyman et al., 1982). Kd is an especially important parameter at the Allen site because for most of the
COI the bulk of the chemical mass in the soil is associated with the solid phase (i.e., sorbed to soil grains
21

rather than dissolved in pore water). In effect, the solid fraction of the soil matrix acts as a large "storage
reservoir" for chemical mass when Kd is large [e.g., metals, many chlorinated solvents, and highly -
chlorinated polycyclic aromatic hydrocarbon (PAH) compounds associated with coal tars and wood -
treating fluids]. Kd is also a very important chemical transport parameter which is used to compute the
chemical retardation factor, Rd, assuming linear equilibrium partitioning of mass between the soil (solid)
and pore -water phases (Hemond and Fechner, 1994):
Rd = 1 + Pb Kd l n,
where Pb is the soil matrix bulk dry density and ne is the effective soil porosity. For example, the
chemical migration rate is directly proportional to hydraulic conductivity and inversely proportional to Rd .
The total contaminant mass in an aquifer is also directly proportional to Rd, as well as aquifer cleanup
times once the source is removed (e.g., Zheng et al., 1991).
Accordingly, it is very important to use accurate Kd values in the CAP Closure Scenario modeling.
Specifically, the CAP Part 1 transport modeling used Kd values that are typically factors of 10 - 100 (i.e.,
one to two orders of magnitude) smaller than the measured site -specific Kd's reported in CAP 1 Appendix
D. In contrast, the CAP Part 2 transport modeling used Kd values that are generally a factor of about 10
larger than the CAP 1 values; however, the CAP 2 Kd's are still on the order of 10 times smaller than the
measured site -specific Kd's reported in CAP 1 Appendix D and CAP 2 Appendix C. Note that the Kd
values in CAP 1 Appendix D are also similar in magnitude to EPA measurements at the Allen Site (EPA,
1985, Table 5.4). Further, soil -water partition coefficients for the CAP Parts 1 and 2 models are much
smaller than most values presented in the literature for the COI (e.g., EPRI, 1984; Baes and Sharp,
1983). This means that, using the actual measured Kd's for the Allen site, the times required to reach
North Carolina water quality standards at the Compliance Boundary are at least a factor of 10 longer (see
additional discussion below) than cleanup times predicted by the CAP Parts 1 and 2 transport models for
many COI.
The CAP 1 modeling report (CAP 1 Appendix C; Section 4.8) argues that the major Kd reductions were
needed due to the following (a similar statement is made in Section 4.8 of CAP 2 Appendix B):
"The conceptual transport model specifies that COls enter the model from the shallow saturated source
zones in the ash basins. When the measured Kd values are applied in the numerical model to COls
migrating from the source zones, some COls do not reach the downgradient observation wells where they
were observed in June/July 2015 at the end of the simulation period. The most appropriate method to
calibrate the transport model in this case is to lower the Kd values until an acceptable agreement
between measured and modeled concentrations is achieved. Thus, an effective Kd value results that
likely represents the combined result of intermittent activities over the service life of the ash basin. These
may include pond dredging, dewatering for dike construction, and ash grading and placement. This
22

approach is expected to produce conservative results, as sorbed constituent mass is released and
transported downgradient."
Considering the approach that was used to develop the chemical transport model (history matching), it is
not true that "the most appropriate method to calibrate the transport model is to lower the Kd values." The
CAP Parts 1 and 2 transport models used an incorrect value (2.65 g/cm3) for the bulk density of
overburden materials; this value is the density of a solid mass of mineral (e.g., quartz) with zero porosity.
The bulk density should have been computed using the total porosity (n) values in CSA Table 11-1 using
the following formula (e.g., Baes and Sharp, 1983):
Pb = 2.65 (1— n)
Based on the Table 11-1 values Pb -- 1.1 - 1.8 g/cm3, which means that the Rd values for the CAP 1 and
2 models were as much as a factor of 2.4 (2.65/1.1) too high before HDR adjusted the Kd values during
calibration. Also, as discussed earlier, the overburden slug test values were about 25 percent too low
due to HDR's data analysis errors. Both of these errors (sorption rate and hydraulic conductivity) resulted
in a modeled transport rate that was up to three times too low before calibration simply due to data input
errors.
At least two other important factors were not considered during the CAP 1 and 2 transport model
calibrations. First, the groundwater flow models are based on average hydraulic conductivity (K) values
within a material zone, but K distributions in aquifers are highly variable (e.g., varying by factors of 3-10,
or more, over distances as small as a few feet: Gelhar, 1984, 1986, 1987; Gelhar and Axness, 1983;
Rehfeldt et al., 1992; Rehfeldt and Gelhar, 1992; Molz, 2015). The Allen site hydrogeology certainly
qualifies as "heterogeneous". This is very important to consider for the CAP transport model calibrations
because it is the high -permeability zones and/or layers that control the time required (Ttravet ) for a
constituent to reach a downgradient observation point, and HDR used differences in observed versus
simulated Tt,,,,et (i.e., time to travel from sources zones to downgradient monitoring wells) as the
justification for lowering measured Kd values.
Second, the history matching that HDR performed is very sensitive to the assumed time at which the
source (i.e., coal ash) is "turned on" and to the assumed distribution of source concentrations (fixed pore
water concentrations) in source area cells. Section 5.2 of CAP 1 Appendix C explains that the source
was activated 58 years ago and that:
"a source term matching the pore water concentrations for each COI was applied within the inactive ash
basin, active ash basin and the ash storage areas at the start of the calibration period. The source
23

concentrations were adjusted to match measured values in the downgradient monitoring wells that had
exceedances of the 2L Standard for each COI in June 2015."
For several reasons it is a major simplification (and generally inaccurate) to use 2015 ash pore water
concentrations to define year-1957 source zone (fixed concentration) boundary conditions. These
reasons include: coal ash was gradually and nonuniformly distributed (spatially and temporally) in ash
basins throughout the 58-year simulation period (not instantaneously in 1957); it is very difficult (or not
possible) to accurately extrapolate geochemical or ash -water leaching conditions (i.e., predict COI pore -
water concentrations) that existed during the 2015 sampling round to conditions that may have existed in
1957 and thereafter; the actual source -area concentration distributions are highly nonuniform, but it is not
clear from the CAP modeling reports how "... source concentrations were adjusted to match measured
values ... ", or if the source area concentrations were nonuniform. All of these uncertainties are further
magnified when using history matching to calibrate a chemical transport model.
Based on the above model input errors and major uncertainties in hydraulic -conductivity variations and
source -term modeling, it is incorrect for HDR to simply reduce Kd values by factors of 10 to 100 below site
measurements (and the large database of literature Kd values) based only on the transport model "history
matching" exercises that HDR performed. My additional comments on the CAP Parts 1 and 2 transport
modeling of Closure Scenarios are listed in the following section.
Geochemical Modeling and Evaluation of Monitored Natural Attenuation
The CAP Part 2 geochemical modeling results do not include quantitative analyses of COI attenuation
rates at the Allen site and are only qualitative in nature. In addition, HDR did not incorporate any
source/sink (e.g., precipitation/dissolution) terms representing geochemical reaction mechanisms in the
CAP 2 chemical transport model to evaluate whether such reactions are important compared to
groundwater concentration changes caused by advection, dispersion, and soil -water partitioning. In this
regard, HDR states in Section 2.10 of CAP 2 Appendix B : "A physical -type modeling approach was
used, as site -specific geochemical conditions are not understood or characterized at the scale and extent
required for inclusion in the model." Indeed, the Electric Power Research Institute (e.g., EPRI, 1984;
page S-8) has extensively reviewed subsurface chemical attenuation mechanisms applicable to the "utility
waste environment" and concluded: (i) precipitation/dissolution has not been adequately studied; and (ii)
"Quantitative predictions of chemical attenuation rates based upon mineralogy and groundwater
composition cannot be made because only descriptive and qualitative information are available for
adsorption/desorption mechanisms."
Nonetheless, HDR performed the geochemical modeling to evaluate the technical basis for its MNA
analysis; however, any quantitative MNA analysis must compare mass transport rates and changes (e.g.,
24

grams/year per unit area normal to a groundwater pathline) in the aquifer for the various active transport
mechanisms in order to determine whether MNA is a viable alternative (e.g., produces meaningful
groundwater concentration reductions) at the Allen site. In Section 6.3.2 of the CAP 2 report HDR
acknowledges that these quantitative evaluations were not performed in CAP 2 and indicated that they
would need to be completed as part of a Tier III MNA assessment. Nevertheless, HDR suggested in the
CAP 2 report that COI concentrations "will" or "may" attenuate over time without completing the
necessary evaluations to reach these conclusions. HDR also states in CAP 2 Section 6.3.4 that "Data
collected to date show that COls will attenuate over time to restore groundwater quality at the Allen site."
HDR also claims in CAP 2, Section 6.3.3 that "the groundwater model did not allow for removal of COI via
co -precipitation with iron oxides, which likely resulted in an over -prediction of COI transport. Completion
of the Tier 11 assessment described in Appendix H has addressed this issue." Finally, HDR concludes in
Section 7.2.2.1 of the CAP 2 report the following regarding the MNA assessment and Appendix H: "The
most significant finding was that the precipitation of iron and manganese serves to remove other COls
through co -precipitation and adsorption, thus confirming that attenuation is occurring." I saw no
quantitative analyses or evidence in the CAP 2 report or related appendices to support these claims. In
fact, the CAP 2 Appendix H emphasizes that much more geochemical data need to be collected and
chemical transport modeling with a source/sink term must be performed in a Tier III assessment to further
assess whether MNA is a viable remedial alternative.
Therefore, the CAP 2 report fails to provide any quantitative evidence supporting COI attenuation due to
co -precipitation with iron or manganese. The second component of COI attenuation evaluated in
Appendix H is chemical sorption to soil. It is important to note that, although the CAP models did not
incorporate a mechanism for co -precipitation with iron or manganese (or any COI sink term), the CAP
models did simulate attenuation due to sorption. Even with the sorption attenuation mechanism included,
CAP 2 Table 4-1 shows that for both the "existing conditions" and "cap -in -place" scenarios the following
COI will exceed North Carolina groundwater standards at the Compliance Boundary 100 years into the
future: antimony, boron, chromium, hexavalent chromium, cobalt, and vanadium. Further, my Table 1
shows that groundwater standards are currently exceeded at the Compliance Boundary for iron,
manganese, and total dissolved solids (i.e., iron, manganese, and TDS contaminant plumes originating in
the source areas have already reached the Compliance Boundary). The conclusions of the MNA Tier I
analyses (CAP 2 Appendix H, page 16) were that antimony, boron, cadmium, chromium, cobalt,
manganese, and nickel show limited evidence of attenuation and should not be evaluated further for
MNA. These Appendix H conclusions are consistent with the Table 4-1 results for these COI. In addition,
my Table 1 results support the Appendix H conclusions that boron, chromium, cobalt, and manganese
are not candidates for MNA because these contaminant plumes have also already reached the
Compliance Boundary. All of these data and CAP 2 modeling results strongly contradict the CAP 2
conclusion (e.g., Section 7.2.2.1) that ".... MNA is the recommended corrective action for the Allen site."
25

Simulation of Closure Scenarios
As discussed below, CAP 1 Closure Scenario simulations greatly underestimate (by factors of 10 or
more) the time frames required to achieve meaningful groundwater concentration reductions in response
to remedial actions. Compared to the Cap -in -Place (CIP) remedial alternative evaluated in the CAP Part
1, the Excavation Scenario results in COI concentration reductions at the Compliance Boundary that are
generally two to four times (2 - 4x) greater compared to Cap -in -Place and best reduces impacts to surface
water. In addition, the time frames to achieve equivalent concentration reductions are factors of 2.5 to 5
shorter for excavation compared to cap -in -place.
Although the CAP 1 modeling showed that Source Excavation outperforms CIP, the CAP 2 modeling did
not simulate an Excavation closure scenario. Nonetheless, the following comparisons between CIP and
Excavation impacts on groundwater concentrations are valid for both the CAP 1 and 2 model results.
This is because the main difference with the CAP 2 transport model (compared to CAP 1) is that
concentration changes resulting from either CIP or Excavation (if it was evaluated in CAP 2) occur much
more slowly (i.e., — 10x slower) in the CAP 2 model due to the much larger Kd (and Rd) values. The CAP
2 transport model also assumed uniform initial COI concentrations equal to HDR's proposed provisional
background concentrations (PPBC), even though the PPBC exaggerate background levels (see above
discussion) and there are no data to suggest that background concentrations should be spatially uniform.
Despite these changes in the CAP 1 and 2 models, the relative differences in groundwater concentrations
between the two closure scenarios remain about the same if the uniform starting (PPBC) COI
concentrations are subtracted from the simulated concentration versus time curves. For these reasons
the following discussions focus on the CAP 1 modeling results.
Source Concentrations for Cap -in -Place Scenario
In this scenario the CAP 1 flow model predicts a cap -induced water -table decline equal to approximately
26 feet (relative to the Existing Conditions simulation) in the center of the inactive ash basin (24 feet in
CAP 2 flow modeling); the predicted water -table decline in the active ash basin is about 36 feet (32 feet
in CAP 2). The geologic cross -sections presented in the CSA show that the saturated coal ash thickness
at several locations is up to twice the value assumed in the Cap -in -Place simulation. For example, the
CSA geologic cross -sections show saturated coal ash thicknesses in the active ash basin equal to about
54 feet at boring AB-25 and in cross-section E-E ; 95 feet across cross-section B-B ; 47 feet at boring
AB-21; and 70 feet across section F-F . Saturated coal ash thicknesses in the inactive ash basin are
equal to approximately 43 feet at boring AB-29; 29 feet at boring AB-29; and 29 feet in cross-section D-
D.
26

This means that under the simulated Cap -In -Place Scenario up to one-half of the coal ash, which is the
source of dissolved COI, would remain saturated and continue to leach constituents into groundwater.
The CAP 1 simulations ignored this fact and set all source concentrations equal to zero (i.e., assumed all
coal ash was dewatered). Therefore, the simulated Cap -in -Place concentrations should be much higher
than the values presented in the CAP.
It is important to note, however, that the CAP groundwater flow model simulations exaggerate the
hydraulic effects of the cap (i.e., overstates water table lowering) due to the no -flow boundaries and the
incorrect (i.e., significantly too low) bedrock permeability. The no -flow boundary conditions along the
entire western, southern, and northern grid boundaries prevent flow into the ash basin when large,
laterally inward hydraulic gradients are created by capping. In addition, the flow model uses a hydraulic
conductivity value that is on the order of 50-100 (or greater) times too small, thus restricting upward flow
from bedrock into the capped area and exaggerating predicted water table lowering.
In addition, a site -specific distribution of groundwater recharge values should have been developed for
this and the other simulation scenarios to take into account site -specific topography and soil types (e.g.,
runoff estimation) and climate data (precipitation, evapotranspiration, etc.; e.g., using the U.S. Army
Corps of Engineers HELP Model; Schroeder et al., 1994). The CAP 1 flow model uses an assumed
uniform value of 5 inches year outside of the Ash Basin System and 8.5 inches/year inside ash basins
(except zero recharge for the Retired Ash Basin) even though the actual values are highly variable across
the Ash Basin System and site land surface. Further, as discussed above, HDR should have used a
leaky -type boundary condition to model ponded areas of the active ash basin. The predicted water table
lowering due to capping is very sensitive to the model recharge value, so more effort should have been
made to develop a site -specific recharge -rate distribution.
Slow and Multirate Nonequilibrium Desorption of COI
Since the 1980's the groundwater industry has learned how difficult it is to achieve water quality
standards at remediation sites without using robust corrective actions such as source removal (Hadley
and Newell, 2012, 2014; Siegel, 2014). Two of the key reasons for this in aqueous -phase contaminated
soil are inherently low groundwater or remediation fluid flushing rates in low -permeability zones and slow,
nonequilibrium chemical desorption from the soil matrix (Culver et al., 1997, 2000; Zheng et al., 2010). A
good example of this is the "tailing effect" (i.e., very slow concentration reduction with time) that is
commonly observed with pump -and -treat, hydraulic containment systems. These factors are also related
to the "rebound effect" in which groundwater concentrations sometimes increase shortly after a
remediation system is turned off (Sudicky and Illman, 2011; Hadley and Newell, 2014; Culver et al.,
1997).
27

The CAP 1 and 2 flow models use different permeability (K) zones, but the scale of these zones is very
large and within each zone K is homogeneous even though large hydraulic conductivity variations (e.g.,
lognormal distribution) are known to exist at any field site over relatively small length scales (Molz, 2015).
Moreover, the CAP transport models assume linear, equilibrium soil -water partitioning which corresponds
to instantaneous COI release into flowing groundwater. The transport code (MT3D) has the capability of
simulating single -rate nonequilibrium sorption, but the Close Scenario simulations did not utilize this
modeling feature. Slow desorption of COI can also be expected at the Allen site because sorption rates
are generally highly variable, and multi -rate (Culver et al., 1997, 2000; Zheng et al., 2010), and Kd
values are nonuniform spatially (Baes and Sharp, 1983; EPRI, 1984; De Wit et al., 1995). The CAP flow
and transport models can be expected to significantly underestimate cleanup times required to meet
groundwater standards at the compliance boundary because they do not incorporate these important
physical mechanisms.
Adequacy of the Kd Model for Transport Simulation
The laboratory column experiment effluent data (CAP 1 Appendix D) generally gave very poor matches
with the analytical (one-dimensional) transport model used to compute Kd values. Since the CAP
transport model solves the same governing equations in three dimensions, the adequacy of the Kd
modeling approach for long-term remedial simulations should have been evaluated in much more detail in
the modeling appendix. The sensitivity analyses that were performed only varied Kd by +/- 20 percent
(Table 7 in CAP 1 Appendix C), which is much too small to address the actual uncertainty.
The transport modeling also did not evaluate alternative nonlinear sorption models such as the Freundlich
and Langmuir isotherms (Hemond and Fechner, 1994), which are input options in the MT31D transport
code. Several of the batch equilibrium sorption experiments (CAP 1 Appendix D) exhibited nonlinear
behavior, and such behavior is commonly observed in other studies (e.g., EPRI, 1984). However, HDR
only computed linear sorption coefficients (i.e., Kd) for the Allen site in CAP Part 1. In CAP Part 2 HDR
did fit Freundlich isotherms to the batch sorption data for selected COI (CAP 2 Appendix C, Tables 1-8)
but did not use these Freundlich isotherm results in the CAP 2 transport modeling. De Wit et al. (1995)
showed that the nonlinear sorption mechanism is similar in importance to aquifer heterogeneities in
extending remediation time frames.
Closure Scenario Time Frames
As outlined in my report, the CAP Part 1 chemical transport model underestimates the time intervals
required to achieve groundwater concentration reductions (i.e., achieve groundwater quality restoration)
by a factor of 10 or more. In other words, the CAP 1 transport model significantly overestimates the rate
at which concentrations may reduce in response to remedial actions such as capping or source removal.
This is due to several factors, including major errors in model input data, model calibration mistakes, field
28

data analysis errors, and oversimplified model representation of field conditions (e.g., hydraulic
conductivity) and transport mechanisms (e.g., chemical sorption/desorption). These limitations of
transport models for realistically predicting cleanup times have been recognized by the groundwater
industry for the past few decades based on hands-on experience at hundreds of extensively -monitored
remediation sites.
Even if we ignore the factors of 10 or more errors in cleanup time predictions with the CAP 1 model, the
remediation time frames for the Excavation Scenarios are still more than two centuries for several
constituents due to slow groundwater flushing rates from secondary sources (surrounding residual soil)
left in place after excavation and due to high chemical retardation factors for most of the COI. However,
excavation of secondary -source material would further accelerate cleanup rates under this alternative.
The CAP 1 simulated Cap -In -Place concentration reduction rates are much slower, compared to
excavation, but are also incorrect (i.e., overestimated) because the cap -induced water -table lowering is
insufficient to dewater all of the source -area coal ash, as discussed above, and the CAP 1 and 2 flow
models overestimate cap -induced water -table lowering due to boundary condition errors. Furthermore,
these simulation times are well beyond the prediction capabilities of any chemical transport model for a
complex field site (especially one that is as geochemically complex as the Allen site). The historical
model -calibration dataset (1957-2015) is also significantly smaller than the predictive (remediation) time
frames. In addition, the "history matching" technique used to calibrate the transport model (e.g., major
reduction in measured Kd values) was not performed correctly by HDR.
Cap -In -Place versus Excavation Closure Scenarios
Although the CAP 1 model underestimates remediation time frames, the CAP 1 Closure Scenario
simulations demonstrate several significant advantages of excavation for restoring site groundwater
quality versus cap -in -place. First, predicted COI concentration reductions in groundwater downgradient
from the ash basin system are generally factors of 2-4 greater with excavation compared to CIP (e.g.,
refer to most of the simulated concentration versus time curves in CAP 1 Appendix C). Further, if HDR
had correctly performed the cap -in -place simulations the predicted CIP concentrations would be much
higher because predicted water -table lowering due to the cap would be insufficient to dewater all of the
coal ash. Second, North Carolina 2L or IMAC standards for many COI are not achieved by CIP but are
achieved by excavation (e.g., CAP 1 Appendix C, Figures 16, 17, 31, 74, and 104-107). Third, the time
frames to achieve equivalent concentration reductions are factors of 2.5 to 5 (2.5 - 5x) shorter for
excavation compared to cap -in -place [e.g., Antimony, 2.5x (Fig. 17), 4x (Fig. 16); Arsenic, 5x (Fig. 30);
Barium, 5x (Fig. 44); Chromium, 4x (Fig. 74); Cobalt, 3x (Fig. 105); Selenium, 2.5 (Fig. 121) ].
Even though the CAP 1 modeling demonstrated that the CIP closure alternative would be much less
effective than excavation, and that CIP would only dewater about one-half of the saturated coal -ash
29

thickness in many areas, HDR eliminated excavation from consideration in CAP 2. In Section 7.1 of the
CAP 2 report HDR assumes that "groundwater flow and geochemical modeling indicates that attenuation
by a combination of sorption, chemical precipitation, and dilution effectively dissipates COls in
groundwater in the vicinity of the ash basins" and that "... it is reasonable to assume that COls remaining
in groundwater will decrease in concentration over time as upgradient non -impacted water moves through
the aquifer." As discussed above, HDR provided no quantitative analysis or evidence in the CAP 2 report
or related appendices to support this claim. Considering that up to one-half of the coal -ash source
material would remain saturated with CIP and that multiple exceedances of groundwater standards at the
Compliance Boundary currently exist (with no historical data to indicate that these Compliance Boundary
concentrations are decreasing with time), it is not reasonable to make sweeping assumptions about future
concentration changes. Tier III MNA analyses require rigorous quantitative evaluations using the CAP
transport model with a source/sink term that incorporates geochemical reactions to support MNA as a
viable corrective action. CAP 2 did not provide this information.
HDR also incorrectly concluded in the recommendations section of CAP 2 Appendix G that "... there are
only two locations where groundwater that has been reported to exceed 2L standards fingers out to the
compliance boundary." I showed above (and in Table 1) that groundwater standards have been
exceeded at multiple locations around the Compliance Boundary.
As discussed above, the CAP Part 2 flow model did include a small number of residential wells (44 of the
219 neighboring private wells and none of the four public water supply wells) located between Southpoint
Road and the active ash basin, but the CAP 2 modeling report (CAP 2, Appendix B) did not show
simulated hydraulic head maps with these residential wells pumping and did not provide any discussion
or analyses of the long-term potential for these wells to capture COI dissolved in groundwater. In CAP 2
section 4.1.5 HDR discusses that fact that the CAP 2 flow model was used to compute 1-year, reverse
particle pathlines for these bedrock residential wells (Figure 16 in CAP 2 Appendix B) to determine their
short-term groundwater capture zones. However, the residential well reverse pathline tracing should
have been performed for a much longer time period (e.g., from 1957 to the present) to evaluate whether
COI may have migrated from source areas to these wells. In addition, if HDR had extended the CAP 2
model grid much farther to the west, south, north, and east the capture zones for the remaining 175
private and four public water supply wells could have been determined, as I discuss earlier in my report.
The CAP Closure Scenarios do not include hydraulic containment remedial alternatives (e.g., gradient
reversal) for the bedrock aquifer that would address the risk of off -site COI transport. As discussed
above, the CSA data show many exceedances of groundwater standards in bedrock not only at the
compliance boundary but also inside the CB. In addition, strong downward groundwater flow components
from the deep overburden to bedrock aquifers were measured during the CSA at multiple locations
30

across the site (and at all but two deep/bedrock well clusters in CAP 2), including the western shoreline of
the Catawba River. The cap -in -place alternative does not address either concentration reduction or off -
site chemical migration control in the fractured bedrock aquifer.
The CAP Parts 1 and 2 do not assess whether water quality standards will be achieved in the tributaries
and wetlands between the ash basins and the Catawba River under any closure scenario. Tributaries
downgradient of the active ash basin currently receive contaminated groundwater discharge from the
active ash basin at seep locations S-3 and S-4. As discussed above, for the cap -in -place scenario a
significant fraction of the source material will remain saturated and dissolved COI will continue to migrate
with groundwater toward these seep locations. Although unaddressed by the model, COI concentration
decreases in groundwater and unsaturated zone pore water due to source removal would also reduce
impacts to tributaries and wetlands that are influenced by the ash basins.
Conclusions
Based on my technical review and analyses of the referenced information for the Allen site I have
reached the following conclusions:
• A total of 44 Compliance Boundary groundwater samples exceeded North Carolina groundwater
standards for these COI: boron, chromium, cobalt, iron, manganese, sulfate, total dissolved
solids, and vanadium. Of these 44 exceedances, 29 were greater than the proposed provisional
background concentrations by HDR, which exaggerate background levels;
• Most of the background wells are either likely to be downgradient from coal -ash source areas, or
appear to be downgradient from coal ash. As a result, groundwater concentration data from
these wells overestimate actual site background levels;
• The statistical analyses of background groundwater concentrations at the Allen site (well AB-1 R)
are invalid. The time periods between groundwater sample collection from this well are too small
and the concentration data are not independent;
• There is a significant risk of chemical migration from the ash basin to neighboring private and
public water supply wells in fractured bedrock. The design of the CAP Parts 1 and 2 flow and
transport models prevents the potential for off -site migration from being evaluated;
• The limited CAP model (Parts 1 and 2) domain size; the no -flow boundary conditions along the
western, southern, and northern boundaries; the incorrect hydraulic conductivity value for the
bedrock aquifer; and incorrect hydraulic boundary condition representations of the Catawba River
and the active ash basin prevent simulation and analysis of off -site COI migration;
31

• The CAP Closure Scenario simulations greatly underestimate (by factors of 10 or more) the time
frames required to achieve meaningful groundwater concentration reductions in response to
remedial actions. This is due to oversimplification of field fate and transport mechanisms in the
CAP model and several model input errors;
• The simulated water table lowering for the Cap -in -Place Scenario is almost a factor of two too low
at several locations in the ash basin system in order to dewater all source material; and the
actual cap -induced water table elevation reduction would be much less than predicted due to the
incorrect no -flow boundary conditions and bedrock permeability in the flow model. Therefore, the
remediation time frames for this scenario would be much greater because part of the source zone
would still be active with the cap installed;
• For either the Existing Condition or Cap -in -Place Model Scenario groundwater concentrations of
coal -ash constituents much higher than background levels will continue to exceed North Carolina
groundwater standards at the Compliance Boundary because saturated coal -ash material and
secondary sources will remain in place;
• Source -area mass removal included in the Excavation Scenario results in COI concentration
reductions at the Compliance Boundary that are generally two to four times greater compared to
Cap -in -Place and best reduces impacts to surface water. In addition, the time frames to achieve
equivalent concentration reductions are factors of 2.5 to 5 shorter for excavation compared to
cap -in -place, and source removal reduces the number of COI that will exceed North Carolina
groundwater standards in the future. Additional excavation of secondary sources would further
accelerate concentration reductions;
• Due to an incorrect boundary -condition representation of the active ash basin, the CAP models
underestimate by more than a factor of two both the mass loading of COI into the Catawba River
and the corresponding Catawba River water concentrations (attributable to coal ash ponds)
estimated by the groundwater/surface-water mixing model;
• The CAP Part 2 geochemical modeling and monitored natural attenuation (MNA) evaluations do
not provide the required quantitative analyses (e.g., numerical transport modeling) of COI
attenuation rates necessary to support MNA as a viable corrective action and are only qualitative
in nature. The CAP 2 chemical transport modeling, which included attenuation by sorption,
demonstrated that MNA is not an effective remedial option for several COI (e.g., antimony, boron,
chromium, hexavalent chromium, cobalt, and vanadium);
• The CAP Closure Scenarios do not include hydraulic containment remedial alternatives for the
bedrock aquifer and do not address the risk of off -site COI transport. CSA data show multiple
exceedances of groundwater standards in bedrock not only at the compliance boundary but also
inside the CB. The cap -in -place alternative does not address either concentration reduction or
off -site chemical migration control in the fractured bedrock aquifer; and
32

• Future Compliance Monitoring at the site should include much more closely -spaced Compliance
Wells to provide more accurate detection, and the time intervals between sample collection
should be large enough to ensure that the groundwater sample data are statistically independent
to allow accurate interpretation of concentration trends.
33

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36

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37