EPA 600/R-08/093 I September 2008 I www.epa.gov/ada
United States
Environmental Proti
Agency
Field Application of a Permeable
Reactive Barrier for Treatment of
Arsenic in Ground Water
v^w.-'a,"
'JSJlitll tVf
*J1UW 'V'P •T'srM.V
sss •
Office of Research and Development
National Risk Management Research Laboratory, Ada, Oklahi
-------
-------
Field Application of a Permeable
Reactive Barrier for Treatment of
Arsenic in Ground Water
Richard T. Wilkin, Steven D. Acree,
Douglas G. Beak, Randall R. Ross,
Tony R. Lee, and Cindy J. Paul
Ground Water and Ecosystems Restoration Division
Office of Research and Development
National Risk Management Research Laboratory, Ada, Oklahoma 74820
-------
Notice
The U.S. Environmental Protection Agency through its Office of Research
and Development funded the research described here. It has been subjected to the
Agency's peer and administrative review and has been approved for publication as an
EPA document. Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.
All research projects making conclusions or recommendations based on en-
vironmentally related measurements and funded by the Environmental Protection
Agency are required to participate in the Agency Quality Assurance Program. This
project was conducted under an approved Quality Assurance Project Plan. The
procedures specified in this plan were used without exception. Information on the
plan and documentation of the quality assurance activities and results are available
from the Principal Investigator.
-------
Foreword
The U.S. Environmental Protection Agency (EPA) is charged by Congress with protecting the Nation's land, air, and
water resources. Under a mandate of national environmental laws, the Agency strives to formulate and implement
actions leading to a compatible balance between human activities and the ability of natural systems to support and
nurture life. To meet this mandate, EPA's research program is providing data and technical support for solving
environmental problems today and building a science knowledge base necessary to manage our ecological resources
wisely, understand how pollutants affect our health, and prevent or reduce environmental risks in the future.
The National Risk Management Research Laboratory (NRMRL) is the Agency's center for investigation of
technological and management approaches for preventing and reducing risks from pollution that threatens human
health and the environment. The focus of the Laboratory's research program is on methods and their cost-
effectiveness for prevention and control of pollution to air, land, water, and subsurface resources; protection of water
quality in public water systems; remediation of contaminated sites, sediments and ground water; prevention and
control of indoor air pollution; and restoration of ecosystems. NRMRL collaborates with both public and private sector
partners to foster technologies that reduce the cost of compliance and to anticipate emerging problems. NRMRL's
research provides solutions to environmental problems by: developing and promoting technologies that protect and
improve the environment; advancing scientific and engineering information to support regulatory and policy decisions;
and providing the technical support and information transfer to ensure implementation of environmental regulations
and strategies at the national, state, and community levels.
This publication has been produced as part of the Laboratory's strategic long-term research plan. It is published
and made available by EPA's Office of Research and Development (ORD) to assist the user community and to
link researchers with their clients. Arsenic is a common ground-water contaminant at hazardous waste sites. The
purpose of this document is to provide a hydrologic and geochemical analysis of a pilot-scale Permeable Reactive
Barrier (PRB) installed to treat ground water contaminated with arsenic. This report will fill a need for a readily
available source of information for site managers and others who are faced with the need to remediate ground
water contaminated with inorganic compounds and are considering the use of this cost-effective technology. The
information provided in this document will be of use to stakeholders such as state and federal regulators, Native
American tribes, consultants, contractors, and other interested parties.
Robert W/Puls, Acting Director
Ground Water and Ecosystems Restoration Division
National Risk Management Research Laboratory
-------
-------
Notice ii
Foreword iii
Figures vi
Tables ix
Acknowledgments x
Abstract xi
1.0 Introduction 1
2.0 Background 2
Arsenic in ground water 2
Arsenic removal from water by zerovaient iron 3
3.0 Site Background 4
4.0 PRB Installation . 6
Construction details 6
5.0 Methodology 8
Well network 8
Water sampling and analysis 9
Arsenic speciation modeling 10
Core sampling and analysis 11
X-ray spectroscopy 12
Hydrologic methods 13
6.0 Results and Discussion 17
Ground-water geochemistry 17
PRB behavior 22
Hydraulic investigation 26
Flux evaluations 34
PRB performance 35
Scanning electron microscopy 38
X-ray absorption spectroscopy 40
7.0 Future Study Improvements 45
8.0 Summary and Relevance to Other Sites 46
9.0 References 48
Appendices 52
Appendix A 52
Appendix B 54
Appendix C 56
Appendix D 57
Appendix E 59
-------
Figure 1. Site aerial photograph showing the ASARCO East Helena smelter, town of
East Helena (MT), primary arsenic source zones, and location of the pilot-scale PRB 4
Figure 2. Installation photos: a) long-arm excavator, b) custom bucket, c) trench with
biopolymcr slum', d) excavated materials, c) bucket with aquifer materials,
f) tremie and trench backfill with granular iron, g) granular iron in super sacks,
and h) construction site. . 7
Figure 3. Aerial photos and map showing: a) arsenic concentrations in ground water
(June 2006), b) locations of pilot-PRB and monitoring wells in the test area, and
c) map showing well locations around and in the PRB 8
Figure 4. Discrete multilevel sampler (DMLS) design and application in monitoring wells 9
Figure 5. Specially designed discrete interval sampler used with a mini bladder pump 10
Figure 6. Photograph of a core segment collected from the PRB after 15 months of
operation (September 2006). 11
Figure 7. Typical equipment used in performance of pneumatic slug tests 14
Figure 8. Schematic diagram of electromagnetic borehole flowmetertest design 15
Figure 9. Modified Durov diagram showing trends in major cations, anions, total
dissolved solids, and arsenic concentrations from the former speiss handling area
to the northern site boundary (data collected June 2006) 17
Figure 10. Long-term trends in the major ion chemistry of ground water collected
from monitoring well PBTW-1. . 18
Figure 11. Geochemical profiles across the saturated aquifer in selected wells:
a) PBTW-1, b) PBTW-2, c) EPA04, and d) DH-50 (see Figure 3 for well locations). . 19
Figure 12. Comparison of total dissolved arsenic concentrations and the sum of
arsenate plus arsenite 21
Figure 13. Arsenic concentrations in solutions saturated with As2O- and As2O3 and
site ground water (open red circles) compared to the MCL for arsenic 21
Figure 14. Comparison of total dissolved arsenic and the fraction of total dissolved
arsenic present as arsenite, As (III) 21
Figure 15. Distribution of redox indicators in ground water near the PRB: a) Fe(II)
contoured from data collected in 8/2003, and b) Eh values in ground water in 7/2007 22
Figure 16. Arsenic concentration trends in ground water upgradient from the PRB:
a) total dissolved arsenic as a function of time in ground water from monitoring
wells EPA02 and EPA08, b) fraction of total dissolved arsenic as arsenite, and
c) depth-resolved concentration trends in monitoring well EPA08 24
Figure 17. Plots showing trends with time: a) DOC concentrations, b) arsenic
concentrations, c) pH, and d) Eh for wells sampling ground water from the pilot-PRB 24
Figure 18. Eh-pH diagram for arsenic. Data points show measured pH and Eh for
ground water in the PRB and plume; color code of points shows the measured speciation
of arsenic 25
-------
Figure 19. Comparison of Eh values of ground water measured using a platinum
electrode and calculated using the As(HI)-As(V) redox couple 26
Figure 20. Depth dependent arsenic concentrations in wells EPA08 and EPA09 and
conceptual model of ground-water transport underneath the pilot-PRB. ..................... 27
Figure 21. Ground-water elevations obtained using data from a pressure transducer/data
logger installed in well DH-17 located approximately 30 m upgradient of the PRB. ............ 27
Figure 22. Shallow potentiometric surface interpreted from ground-water elevation
measurements obtained using an electronic water level indicator on April 1, 2008 28
Figure 23. Wells instrumented with pressure transducers/data loggers and used to
characterize hydraulic gradient fluctuations on a daily basis 29
Figure 24. Distribution of the magnitude of the hydraulic gradient near the PRB estimated
on a daily basis using data obtained from pressure transducers/data loggers installed
in wells DH-17, PBTW-2, and EPA06 29
Figure 25. Ground-water monitoring wells within the tracer test area of the PRB 30
Figure 26. Ratio of the measured bromide ion concentration to the concentration in the
injected slug as a function of time. 30
Figure 27. Distance of tracer migration as a function of the time required for the first
arrival at each monitoring location 30
Figure 28. Results of pneumatic slug tests within the PRB using a series of wells each
screening a 0.76 m vertical interval. 32
Figure 29. Results of pneumatic slug tests within the PRB after 5 months and 16 months 32
Figure 30. Hydraulic conductivity distribution estimated for materials adjacent to the
screened intervals of wells DH-17, PBTW-1, and PBTW-2 based on characterization
using borehole flowmeter techniques 33
Figure 31. Hydraulic conductivity distribution estimated for materials adjacent to the
screened intervals of wells EPA02, EPA08, EPA09, and EPA 10 based on characterization
using borehole flowmeter techniques 33
Figure 32. Comparison of hydraulic conductivity distribution adjacent to well EPA08
estimated using borehole flowmeter techniques and the results of slug tests performed
in a series short-screened wells within the PRB 34
Figure 33. Estimation of arsenic flux entering and leaving the PRB as a function of depth. ......... 35
Figure 34. Aragonite saturation indices in ground water as a function of pH, upgradient and
within the PRB 35
Figure 35. Solid-phase concentrations of inorganic carbon, sulfur, and arsenic in PRB
core materials 36
Figure 36. Sulfate removal within the PRB as a function of time and depth. 36
Figure 37. Chromatograph of arsenic speciation for ground water entering the
PRB (well EPA08) and ground water from well TR9 containing thioarsenic species. ........... 37
Figure 38. Solubility of As(III) phases as a function of pH and dissolved sulfide concentration. .... 37
Figure 39. Solubility of As(V) phases as a function of pH. .................................. 37
Figure 40. Concentration of selenium in ground water upgradient, in, and downgradient
of the PRB 38
-------
Figure 41. SEM photomicrographs: A) image of polished thin-section showing development
of corrosion products. B) platy particles formed on the surface of an iron granule with
EDX spectra and elemental composition 39
Figure 42. Normalized XANES spectra for arsenic reference materials used for the
XANES analysis and LCF fitting of unknown samples. 40
Figure 43. Normalized XANES spectra for unknown samples from the ASARCO Smelter site 41
Figure 44. a) Depth-dependent concentrations of total arsenic in aquifer solids retrieved from
the well boring near the location of the PRB, and b) speciation of arsenic expressed as the
fraction of total arsenic as As(V) in the solid samples 42
Figure 45. Powder x-ray diffraction scans of aquifer materials collected from a well boring taken
adjacent to the pilot-PRB.. . 42
Figure 46. Comparison of LCF fitting results using: a) coordinated or b) sorbed reference
materials for PRB core sample Core 1 30-34F 44
Figure 47. Ternary diagram showing the solid-phase arsenic speciation based on XANES LCF
in samples collected from the PRB, source zone, and aquifer adjacent to the pilot-PRB 44
Figure Dl. Hydrograph of ground-water elevations measured at well DH-17 57
Figure D2. Hydrograph of ground-water elevations measured at well EPA02 57
Figure D3. Hydrograph of ground-water elevations measured at well EPA06 57
Figure D4. Hydrograph of ground-water elevations measured at well PBTW-2. . 57
Figure D5. Hydrograph of ground-water elevations measured at well TR8 58
Figure El. Potentiometric surface at the water table in vicinity of PRB on June 19, 2002. .......... 59
Figure E2. Potentiometric surface at the water table in vicinity of PRB on September 26, 2002 59
Figure E3. Potentiometric surface at the watertable in vicinity of PRB on August 14, 2003 60
Figure E4. Potentiometric surface at the watertable in vicinity of PRB on May 31, 2005 60
Figure E5. Potentiometric surface at the watertable in vicinity of PRB on July 20, 2005 61
Figure E6. Potentiometric surface at the watertable in vicinity of PRB on October 6, 2005. ........ 61
Figure E7. Potentiometric surface at the watertable in vicinity of PRB on June 6, 2006. ........... 62
Figure E8. Potentiometric surface at the water table in vicinity of PRB on September 18. 2006. ..... 62
Figure E9. Potentiometric surface at the water table in vicinity of PRB on January 24, 2007 63
Figure E10. Potentiometric surface at the watertable in vicinity of PRB on July 18/19, 2007 63
Figure Ell. Potentiometric surface at the water table in vicinity of PRB on October 1, 2007 64
Figure E12. Potentiometric surface at the watertable in vicinity of PRB on April 1, 2008 64
Figure E13. Potentiometric surface at the watertable in vicinity of PRB on June 24, 2008 65
-------
Tables
Table 1. Samples analyzed using x-ray absorption spectroscopy 12
Table 2. Edge and white-line positions for reference materials used in data analysis
of the PRB cores, source zone materials and well borings 13
Table 3. Comparison of hydraulic gradients calculated using multiple wells in the vicinity
of the PRB and using data from three wells (DH-17, PBTW-2, and EPA08) 14
Table 4. Ground-water parameters in selected wells 23
Table 5. Estimates of hydraulic conductivity in units of m d"1 obtained from pneumatic
slug tests performed in wells located within the PRB 31
Table 6. Results of the LCF fitting of the unknown samples collected from the Asarco smelter. 43
Table Al. Wells used in the design and assessment of the PRB 52
Table Bl. Method reporting limits for selected analytes during five separate PRB sampling
events and results of selected duplicate analyses 54
Table B2. Results of blank tests and pump rinsate results 55
Table Cl. Thermodynamic data for arsenic oxyanions 56
-------
Acknowledgments
Research described in this report was partially funded by a Cooperative Research and Development Agreement
(CRADA, 278-03) between ASARCO, Inc. and the U.S. EPA. The authors would like to acknowledge cooperation
with ASARCO and U.S. EPA Region 8 for the successful completion of this project. Jon Nickel and Robert Miller
from ASARCO are especially thanked for their helpfulness and knowledge shared during many site visits. Linda
Jacobson, Randy Breeden, Chuck Figur, and Scott Brown (U.S. EPA) are thanked for their efforts in seeing this project
through. Susan Zazzali (U.S. EPA) is thanked for initially focusing our attention on the East Helena site. Shaw
Environmental provided support both in the field and laboratory, especially Elaine Coombe, Sujith Kumar, Don Janz,
and Steve Markham. We thank Mary Sue McNeil, Pat Clark, Michael Brooks, Ken Jewell, Chunming Su, Frank Beck,
Linda Callaway, Brad Scroggins, Ralph Ludwig, Robert Puls (U.S. EPA), Hsing-Lung Lien (National University of
Kaohsiung, Taiwan), and Joanne Smieja (Gonzaga University) for discussions and help in the field and lab.
The authors greatly appreciate support provided by the staff of the Dow-Northwestern-DuPont Collaborative Access
Team and the Pacific Northwest Consortium Collaborative Access Team at Argonne National Laboratory. Use of
Argonne's Advanced Photon Source is supported by the U. S. Department of Energy (US DOE), Office of Science,
Office of Basic Energy Sciences, under Contract DE-AC02-06CH11357. PNC-XOR facilities at the Advanced Photon
Source, and research at these facilities, are supported by the US DOE Basic Energy Science, a major facilities access
grant from NSERC, the University of Washington, Simon Fraser University and the Advanced Photon Source. DND-
CAT is supported by E.I. Dupont de Nemours & Co., the Dow Chemical Company and the State of Illinois.
The report was reviewed by Bruce Manning (San Francisco State University), Ralf Kober (University of Kiel),
Kirk Scheckel (U.S. EPA), and Charles Pace (NewFields Companies LLC). Their comments helped improve the
presentation and discussion of the study results.
-------
Abstract
Contamination of ground-water resources by arsenic is a widespread environmental problem; consequently, there
is an escalating need for developments and improvements of remedial technologies to effectively manage arsenic
contamination in ground water and soils. In June 2005, a 9.1 m long, 14 m deep, and 1.8 to 2.4 m wide (in the
direction of ground-water flow) pilot-scale permeable reactive barrier (PRB) was installed at a former metal smelting
facility, located near Helena, Montana. The reactive barrier was designed to treat ground water contaminated with
moderately high concentrations of both arsenite and arsenate. The reactive barrier was installed over a 3-day period
using bio-polymer slurry methods and modified excavating equipment for deep trenching. The reactive medium
was composed entirely of granular iron which was selected based on long-term laboratory column experiments. In
laboratory experiments, arsenic removal by zerovalent iron is controlled by adsorption and co-precipitation with
iron corrosion products. Previous studies indicate removal capacities on the order of 1 to 10 mg arsenic per gram of
granular iron. A monitoring network of approximately 40 ground-water sampling points was installed in July 2005.
Monitoring results indicate arsenic concentrations >25 mg I/1 in wells located hydraulically upgradient of the PRB.
Within the PRB, arsenic concentrations are reduced to 2 to <0.01 mg I/1. After 2 years of operation, monitoring points
located within 1 m of the downgradient edge of the PRB showed significant decreases in arsenic concentrations at
depths intervals impacted by the emplaced zerovalent iron. Arsenic removal in the PRB results from several pathways
involving adsorption to iron oxide and iron sulfide surfaces. These different uptake processes lead to multiple
oxidation states and bonding environments for arsenic in the reactive medium as indicated using spectroscopic
methods. This report covers aspects of site characterization, remedial design and implementation, and monitoring
results for this pilot-scale PRB, including a flux-based analysis for arsenic.
-------
-------
1.0
Introduction
The permeable reactive barrier (PRB) technology has
gained acceptance as an effective passive remediation
strategy for the treatment of a variety of chlorinated
organic and inorganic contaminants in ground water
(e.g., O'Hannesin and Gillham, 1998; Blowes et al.,
2000). The technology combines subsurface fluid-
flow management with contaminant treatment by
combinations of chemical, physical and/or biological
processes. Application of PRBs for treatment of
contaminated ground water has advantages over
traditional pump-and-treat systems in that PRBs are
passive and are expected to require minimal operation
and maintenance expenditures. More than two hundred
implementations of the technology worldwide have
proven that passive reactive barriers can be cost-effective
and efficient approaches to remediate a variety of
hazardous compounds of environmental concern (ITRC,
2005).
Few well documented case studies are available that
evaluate the field performance of these in-situ systems,
especially with respect to the treatment efficiency of a
variety of contaminant types and including examples
from complex hydrogeologic environments. In some
cases, PRB applications for ground-water remediation
have failed to achieve cleanup results as expected from
bench-scale tests. For example, Morrison et al. (2006)
reported on a zerovalent iron system that showed
sooner than expected breakthrough of molybdenum
and uranium. Performance failure was determined to
be related to a sequence of events from the continual
buildup of mineral precipitates on the reactive medium,
loss of pore space, development of preferential flow
paths, and finally to complete bypass of the zerovalent
iron and loss of hydraulic control. Research efforts
over the past decade point to complex behavior in PRB
systems, as biogeochemical processes in the reactive
medium govern contaminant removal and influence
processes that control fluid flow through porous media
(e.g., U.S. EPA, 2003a; Liang et al., 2005; Li et al.,
2006). Clearly these factors need to be better understood
in order to improve the design and implementation of
PRBs for ground-water remediation.
This report presents a pilot-scale examination of the PRB
technology with zerovalent iron for treatment of arsenic
in contaminated ground water. The goals of this report
are to (1) document the design and construction of the
pilot-scale PRB; (2) describe the hydraulic and reactive
performance of the PRB; and (3) document variation in
arsenic behavior and related geochemical factors within
the PRB and the contaminated aquifer system. The
study serves to fill in a needed aspect of the technology
continuum that encompasses bench-scale testing, pilot-
scale field testing, and full-scale field applications.
-------
2.0
Background
Arsenic in Ground Water
Arsenic is a well-known toxic element that the U.S.
Environmental Protection Agency and the World
Health Organization list as a carcinogen. In subsurface
systems, including soils, sediments, and ground water,
arsenic is present in a variety of chemical forms that are
influenced by changes in biogeochemical conditions.
Arsenic occurs in association with minerals, such as
sulfides (e.g., pyrite), metal oxides (e.g., goethite),
clays, silicates, and carbonates. The most important
natural sources of elevated arsenic in ground water
are iron sulfides and iron oxides, partly because of
the abundance of iron-containing minerals in aquifers
(Smedley and Kinniburgh, 2002). Polizzoto et al. (2006)
provide evidence that arsenic present in sulfide minerals
is the dominant source of arsenic in Holocene aquifer
sediments of Bangladesh. There arsenic released from
sulfide minerals in near-surface oxidizing environments
subsequently adheres to iron oxyhydroxides which are
then subject to dissolution in anaerobic environments.
In the anaerobic aquifer, arsenic is not retained by
the aquifer solids but instead remains in the aqueous
phase where tragically it is pumped through wells and
consumed as drinking water by an estimated 57 million
people.
As suggested above in the Bangladesh example, arsenic
exhibits fairly complex chemical behavior in the
environment and may be present in several oxidation
states (-III, 0, III, V). In aquatic environments, two
oxidation states are mainly encountered (Cherry et
al., 1979; Ferguson and Gavis, 1972). The dominant
form in oxic waters is arsenate, an oxyanion with
the +5 oxidation state. Arsenate can be present as
various protonated forms depending on pH: H3AsO4,
H2AsO4", HAsO42", and AsO43". In anoxic waters, the
most common form of arsenic is arsenite, an uncharged
species (below pH 9.2, H3AsO3) with a +3 oxidation
state. Because arsenite is typically uncharged in
ground-water systems, it is usually found to be mobile
in solution. Ferrous iron is able to reduce arsenate to
arsenite in the presence of iron oxyhydroxide surfaces,
but not in homogeneous solution (Johnston and Singer,
2007a). In some ground-water settings oxygen is an
available, thermodynamically favorable oxidant for
arsenite. However, the rate of arsenate formation via
oxidation of arsenite by molecular oxygen is generally
sluggish and highly pH dependent (Cherry et al., 1979).
Toxicological studies show arsenite to be the more
hazardous form of arsenic. Thus, reducing conditions,
which generally favor arsenic mobility, also favor the
formation of the more toxic oxyanion of arsenic. Ground
water can also contain organoarsenic species, such
as monomethylarsenic acid and dimethylarsenic acid
(Cullen and Reimer, 1989). In general, organoarsenic
compounds are less toxic than their corresponding
oxyanions. There are also arsenic-sulfur species
(thioarsenic species) that provide additional complexity
to arsenic speciation in reducing environments. Beak
et al. (2008) indicate that at sulfide concentrations
>100 (oM, thioarsenic species can become dominant over
the oxyanion species, arsenite and arsenate. Conversion
from arsenite to thioarsenite species is believed to reduce
arsenic toxicity (Rader et al., 2004). Under extremely
reducing conditions elemental arsenic and arsine may
be present, although their occurrence in ground water
systems has not been widely documented.
Arsenate, arsenite, and thioarsenic species are highly
soluble anions and will tend to remain in solution after
being released from the mineral-water interface. Indeed,
Magalhaes (2002) points out that the primary challenge
in mitigating arsenic mobility in the environment is tied
to the high solubility of metal arsenites and arsenates.
Under oxic conditions, arsenic can be released to ground
water by dissolution of iron sulfides or by desorption
from iron oxides due to an increase in pH or competition
with other anions (Welch et al., 2000; Smedley and
Kinniburgh, 2002). As pointed out above in the
Bangladesh study, a particular geochemical environment
that favors release of arsenic is the onset of iron-
reducing conditions, which results from the degradation
of organic carbon. Under iron-reducing conditions,
arsenic associated with iron hydroxides, oxyhydroxides,
or oxides can be released to ground water by reductive
desorption or reductive dissolution. Reductive
desorption occurs when arsenate is reduced to arsenite,
which is less strongly sorbed to iron oxides; reductive
dissolution of iron minerals releases arsenic that is part
of the iron-mineral structure or sorbed at the mineral-
water interface. On the other hand, in sulfate-reducing
environments and environments where iron oxides
are stable, iron sulfides and iron oxides are important
sinks for arsenic, so the formation and stability of these
minerals can retard the migration arsenic in ground water
(U.S. EPA, 2007).
In January 2006 the U.S. Environmental Protection
Agency adopted a new maximum contaminant level
(MCL) for arsenic of 0.01 mg I/1, decreased from the
previous level of 0.05 mg I/1. This revision of the MCL
recognizes the detrimental health effects associated with
arsenic in drinking water, including bladder, skin, and
-------
lung cancers, diabetes, and neurological dysfunction
(National Research Council, 1999). Elevated
concentrations of arsenic from natural sources (>0.05
mg I/1) have been widely documented, for example, in
Argentina, Bangladesh, Chile, West Bengal, Mexico,
Taiwan, Mexico, and parts of the United States (e.g.,
Mandal, 1997; Nickson et al., 1998; Welch et al.,
1988; Del Razo et al., 1990; McArthur et al., 2001;
Rahman et al., 2001; Nordstrom, 2002). While arsenic
occurs naturally, it may also be found as a result of a
variety of industrial processes, including mining, metal
refining, manufacture and use of arsenical pesticides
and herbicides, release of industrial effluents, leather
and wood treatments, and chemical waste disposal.
These industrial activities have created a long legacy
of arsenic pollution throughout the United States
where arsenic is a common contaminant of concern
at Superfund and RCRA sites. For example, in 1996
arsenic contamination was found at 226 Superfund sites
(U.S. EPA, 1997). Levels of arsenic in ground water
>1-10 mg I/1 are not unusual at many Superfund and
RCRA sites in the United States. The combination of
high toxicity and widespread occurrence of arsenic has
created a pressing need for the development of arsenic
treatment strategies in ground water. Furthermore, it
is expected that the new MCL for arsenic will impact
cleanup expectations at Superfund and RCRA sites
across the country.
Arsenic Removal from Water by Zerovalent Iron
Lackovic et al. (2000) concluded that zerovalent iron
could be used in PRBs to remove inorganic forms of
arsenic from ground water, including arsenate and
arsenite. These researchers noted that the removal
mechanism for arsenic contrasted with that of
chlorinated hydrocarbons (reductive dechlorination)
and hexavalent chromium (reductive precipitation), and
involved either adsorption or precipitation on the iron
surface. Lackovic et al. (2000) further found that arsenic
removal efficiency improved with time, perhaps related
to corrosion of the zerovalent iron and production of new
sorption sites for arsenic uptake.
Since the Lackovic et al. (2000) study there has been a
considerable research effort focused on zerovalent iron
and its potential for removing arsenic from water (e.g.,
Ramaswami et al., 2001; Su and Puls, 2001a,b; Farrell
et al., 2001; Morrison et al., 2002; Manning et al., 2002;
Melitas et al., 2002; Su and Puls, 2003; Nikolaidis et al.,
2003; Bang et al., 2005a,b; Lien and Wilkin, 2005;
Leupin and Hug, 2005; Kober et al., 2005; Sun et al.,
2006; Yuan and Chiang, 2007; Biterna et al., 2007). A
common finding of these studies is that arsenic removal
from water is attributable to adsorption onto corrosion
products of zerovalent iron, including iron hydroxides,
oxyhydroxides, and mixed valance Fe(II)-Fe(III) green
rusts (Farrell et al., 2001; Melitas et al., 2002; Manning
et al., 2002; Su and Puls, 2003; Leupin and Hug, 2005;
Lien and Wilkin, 2005; Bang et al., 2005b; Yuan and
Chiang, 2007). Batch studies to examine the effects of
anion competition for arsenite and arsenate adsorption
indicate that phosphate causes a significant decrease
in the removal rate of arsenic, followed by silicate,
chromate, molybdate, carbonate, and nitrate (Su and
Puls, 200Ib). Borate and sulfate caused only slight
reductions in arsenic uptake rates. Uptake capacities
determined in controlled laboratory column tests have
ranged from about 1 to 7.5 mg As per g of zerovalent
iron (Su and Puls, 2003; Nikolaidis et al., 2003; Lien and
Wilkin, 2005).
The detailed nature of arsenic uptake mechanisms
onto zerovalent iron has been probed with solid-phase
characterization tools sensitive to arsenic, including
x-ray absorption spectroscopy, Auger electron
spectroscopy, x-ray photoelectron spectroscopy, and wet
chemical extractions (e.g., Su and Puls, 200la; Farrell
et al., 2001; Manning et al., 2002; Melitas et al., 2002;
Nikolaidis et al., 2003; Lien and Wilkin, 2005; Bang et
al., 2005a). Application of these techniques suggests
that several processes may be important during the initial
removal of arsenic from water and during long-term
aging processes, such as adsorption, precipitation, co-
precipitation, and redox transformation. For example,
studies suggest that sorbed As(III) can transform to
As(0) (Bang et al., 2005a) or As(V) (Su and Puls, 2001a;
Manning et al., 2002; Lien and Wilkin, 2005; Leupin and
Hug, 2005) depending on aging conditions, but reduction
of sorbed As(V) to As(0) has not been observed (Farrell
et al., 2001; Bang et al., 2005a). Reduction of As(V)
to As(III) is indicated in some studies (e.g., Su and
Puls, 2001a) but not in others (e.g., Farrell et al., 2001),
possibly due to the variable nature of zerovalent iron and
water chemistries used in laboratory experiments.
Field-based applications of zerovalent iron for arsenic
treatment are few in comparison to laboratory bench
tests (e.g., Morrison et al., 2002; Nikolaidis et al.,
2003; Vlassopoulos et al., 2005; Bain et al., 2006), and
evaluations of uptake mechanisms are not available to
compare with laboratory tests. An additional factor that
needs to be accounted for in field tests is the impact of
microorganisms. Activity of sulfate-reducing bacteria in
zerovalent iron PRBs has been documented (Roh et al.,
2000; Furukawa et al., 2002; Wilkin et al., 2003, 2005).
Production of biotic sulfide adds additional pathways
for removal of inorganic contaminants via adsorption to
and precipitation of insoluble metal sulfide precipitates.
Indeed, several studies suggest that arsenic removal
processes in zerovalent iron are linked to interactions
with sulfur (Ramaswami et al., 2001; Nikolaidis et al.,
2003; Kober et al., 2005). This pilot-scale study allows
for an examination of arsenic uptake processes in a
complex field setting.
-------
3.0
Site Background
The ASARCO East Helena plant was a custom smelter
located just to the south of the City of East Helena,
Montana. The site is located at the southwestern margin
of the Helena Valley. The plant is bounded to the south
by Upper and Lower Lake and, to the east and northeast,
by Prickly Pear Creek (Figure 1). The geology and
hydrogeology of the area have been described in detail
by Lorenz and Swenson (1951) and summarized by
Briar and Madison (1992). Surficial geology in the area
of the site has been mapped by Stickney (1987).
Figure 1. Site aerial photograph showing the ASARCO East Helena smelter, town of East Helena (MT), primary
arsenic source zones, and location of the pilot-scale PRB.
-------
The Helena Valley is an intermontane basin bounded by
sedimentary, metamorphic, and igneous rocks (Briar and
Madison, 1992). The valley is underlain by a sequence
of layered sediments that are not well characterized at
depth. During the Quaternary period of the Cenozoic
era, streams, including Prickly Pear Creek, deposited
sediments in channel-fill and alluvial-plain environments
over much of the central portion of the valley. The
geology of the valley near the site was also affected by
glaciation during the Pleistocene period. Alpine glaciers
at the headwaters of Prickly Pear Creek increased the
coarse sediment load of the creek through increased
stream discharge during spring melting and the draining
of glacial lakes. With respect to the shallow portion
of the valley-fill aquifer system, these processes have
resulted in a complex sequence of stratified lenses of
cobbles, gravel, and sand with interlayered silt and
clay. The sequence grades from predominantly cobbles,
gravel, and coarse sand where streams such as Prickly
Pear Creek enter the valley to predominantly sand, silt
and clay near Lake Helena in the northern part of the
valley.
Surface sediments at much of the plant site have been
mapped by Stickney (1987) as smelter tailings. The
native geologic materials bounding the northern,
southern, and eastern margins of the plant are described
as Holocene-age stream-channel deposits that are
moderately sorted, fine to coarse sandy pebble to cobble
gravel and Holocene terrace and alluvial fan deposits
consisting of moderately sorted pebble to cobble gravel
in a silty sandy matrix. Based on geologic logs produced
from continuous split-spoon sampling at wells PBTW-1
and PBTW-2 in the area of the PRB (Figure 1), the
shallow aquifer consists of relatively coarse-grained
but highly variable, unconsolidated alluvial deposits
containing mixtures of cobbles, gravel, sand with some
silt. Fine-grained material, described as volcanic ash
deposits, underlie the shallow aquifer materials in this
area of the site.
The climate of the Helena Valley is semiarid with
average annual precipitation between 10 and 12 inches
per year in the vicinity of the site (Briar and Madison,
1992). Precipitation is generally highest during the late
spring/summer months and lowest during the fall and
winter months. Prickly Pear Creek, bounding the eastern
portion of the smelter site and supplying water to Upper
Lake at the southern site boundary, is the largest of the
four principal streams flowing into the valley. During
a hydrologic study performed by the U.S. Geological
Survey in 1990/1991 (Briar and Madison, 1992),
streamflow in Prickly Pear Creek was highest during the
months of May and June and lowest during the months
of December and January. The principal sources of
recharge to the Helena Valley aquifer system as inferred
by the 1990/1991 study are infiltration from streams,
infiltration from irrigation-related sources, and inflow
from fractures in surrounding bedrock.
The plant operated for over 100 years starting around
1888. Lead and zinc smelting operations resulted in the
deposition of lead, arsenic, copper, zinc, cadmium, and
other hazardous substances into soil and surface waters
around the plant. Ground water underneath the site is
contaminated in locations with arsenic, selenium, lead,
cadmium, and zinc; plumes of arsenic and selenium
have migrated offsite whereas the occurrence of other
dissolved metals appears to be restricted within site
boundaries. The East Helena Site was listed on the
National Priorities List (NPL) in 1984. ASARCO shut
down plant operations in April 2001 and currently plant
demolitions and remedial investigations are underway.
Arsenic contamination in the ground water stems from
several identified source areas. The primary source
area for arsenic is located near the former speiss
handling area (Figure 1). Speiss is the lightest molten
phase produced in lead smelting operations and is
characteristically enriched in arsenic and sometimes
antimony. Other source areas include Lower Lake and
the former acid plant sediment drying area. Arsenic
concentrations in ground water exceed the 0.01 mg L'1
maximum concentration limit on the plant site and in
an area hydraulically downgradient of the plant site.
The highest concentrations occur in the former speiss
handling area, the acid plant area, and the former acid
plant sediment drying area.
-------
4.0
PRB Installation
The pilot-PRB is a 9.1m long (perpendicular to ground
water flow), 13.7 m deep, and 1.8 to 2.4 m wide
(parallel to ground water flow) installation of granular
zerovalent iron (-8 + 50 mesh, Peerless Metal Powders
and Abrasives, Detroit, MI). The reactive barrier was
installed over a 3-day period using bio-polymer slurry
methods and modified excavating equipment for deep
trenching (see Figure 2 for installation photos). The
reactive medium was composed entirely of granular
iron which was selected based on long-term laboratory
column experiments (Lien and Wilkin, 2005). The
trench, located approximately 280 m hydraulically
downgradient of the speiss handling area, was backfilled
with a 7.6-m thick layer of granular iron (from 13.7 to
6.1m below ground surface) and a 6.1-m thick layer
of sand (from 0 to 6.1 m below ground surface). The
top of the granular iron zone is located >1 m above
the maximum ground water level observed during site
characterization studies. The base of the granular iron
zone is located approximately 1 m above the confining
ash tuff deposit, and therefore the PRB is a "hanging
wall". This configuration was a planned aspect of the
study in order to ensure that the lower confining unit was
not breached during construction of the pilot system,
to minimize costs, and to examine potential by-pass
processes. The PRB contains approximately 174 t of
granular iron with an initial porosity of over 50%.
After completing the backfill of granular iron and sand
there was excess bio-polymer slurry in the trench and
within the pore space of the zerovalent iron medium. In
order to re-establish the permeability of the surrounding
aquifer and to permit ground water to flow through
the PRB, it was necessary to initiate breakdown of the
slurry. This was accomplished by (1) breaking down the
bio-polymer slurry to simple carbohydrates (monomers)
and (2) encouraging native soil microbes to consume the
carbohydrates. Two air lift pumps were set up to extract
slurry from eight temporary wells and discharge the
slurry over the surface of the backfill. The set up of the
airlift pumps permitted slurry to circulate from the wells
through the backfill layers and back to the wells into the
reactive medium. Liquid enzyme breaker was placed
into the temporary wells. The degradation or "breaking"
process of the bio-polymer slurry took 3 d to complete.
A Marsh funnel viscosity of less than 30 seconds
indicated the slurry was broken. The pumping and
slurry recirculation continued until a minimum of
3 pore volumes of the trench was circulated to flush and
develop the trench.
Construction Details
Shaw Environmental, Inc. performed the construction
in accordance with Work Assignment WA-RB-1-8
issued by the EPA under Contract No. 68-C-03-097
(Shaw, 2005a) and described in Shaw (2005b) from
which the following narrative on construction details
is derived. Geo-Solutions was the subcontractor
to Shaw for the bio-polymer slurry portion of the
project. Prior to excavation, asphalt was cut with a
walk behind saw; approximately 13 cubic yards (cy)
of asphalt were disposed of offsite at the local sanitary
landfill. Excavation started on June 4, 2005. The PRB
trench was excavated initially with a Cat 320 smooth
bucket excavator in the utility corridor. Spotters with
shovels were also utilized during the first 1 m of the
excavation to ensure that no underground utilities were
present. A long-arm excavator (Komatsu PC750) with
a ~l-m wide bucket completed the remaining trench
excavation. Excavated soils were placed in a lined
spoils containment area, which was located within the
reach of the long-arm excavator.
Bio-polymer slurry was added to the PRB trench in
order to stabilize the trench walls. Excess water in the
excavated soil was allowed to spill from the excavator
bucket back into the trench before unloading the soil into
a lined spoils containment area. Measurement below
the bio-polymer slurry level was made with a sounding
cable. The slurry consisted of guar gum (Rantec G150),
water, and additives. A 20,000 gallon frac tank was used
to temporarily store the mixed slurry. A smaller tank
was used to hold potable water prior to mixing.
Zerovalent iron was placed into the trench using tremie
equipment. The tremie method was used to minimize
the potential for segregation and to ensure adequate
backfill density. The tremie consisted of jointed
vertical pipe connected to a hopper that had legs that
straddled the trench. Granular iron was backfilled into
the trench from the bottom or 13.7 m below ground
surface to 6.1 m below ground surface. The granular
iron was delivered in "super sacks" that weighed about
3,000 pounds each. Prior to installing the PRB, the
super sacks were stored at an ASARCO warehouse. The
super sacks were transported to the PRB location. The
granular iron was moisture conditioned and mixed with
bio-polymer slurry in a bedding box prior to backfilling.
The conditioned zerovalent iron was placed in the tremie
hopper where it fell through a tremie pipe and flowed
out on to the bottom of the trench. It was necessary
to move the tremie laterally three times and raise the
-------
tremie equipment vertically when the tremie pipe was
filled in order to continue the backfill process. A total of
116 iron-filled super sacks was used to backfill the PRB
trench.
Sand material, which consisted of coarse bedding sand,
was then placed into the trench on top of the zerovalent
iron. The installation work plan called for backfilling
the sand layer using the tremie system (Shaw, 2005a)
to avoid segregation issues. However, the sand was
backfilled using a loader bucket because segregation
was not an issue with the coarse bedding sand used. The
sand material was backfilled to near ground surface.
Approximately 228 tons of coarse bedding sand was
used.
Figure 2. Installation photos: a) long-arm excavator, b) custom bucket, c) trench with biopolymer slurry,
d) excavated materials, e) bucket with aquifer materials, f) tremie and trench backfill with granular iron,
g) granular iron in super sacks, and h) construction site
-------
5.0
Methodology
Well Network
Ground-water monitoring wells were installed in a
selected region of the site for site characterization
purposes prior to the installation of the pilot-PRB.
Subsequently, additional monitoring wells were installed
upgradient, within, and downgradient of the PRB
following installation to monitor reactive and hydraulic
performance of the pilot system (Figure 3). Wells
within the PRB were installed using direct push methods
(Geoprobe Systems). Specifically, the wells were
installed through steel rods allowing the iron to collapse
around the well as the rods were removed. Most of
the wells within the PRB are constructed using either
2.54 cm (1 in) or 5.08 cm (2 in) schedule 40 PVC casing
and screen with a slot size of 0.051 cm (0.020 in), and
are completed between 7.6m and 14.6m below ground
surface. In addition, five wells were installed using
3.175 cm (1.25 in) OD three-channel tubing. In other
wells screened across the entire saturated zone, detailed
concentration and geochemical profiles were obtained
\
3m
c)
i Multiport well
Figure 3. Aerial photos and map showing: a) arsenic concentrations in ground water (June 2006), b) locations of
pilot-PRB and monitoring wells in the test area, and c) map showing well locations around and in the
PRB.
-------
Discrete multi-
level sampler (DMLS)
Well casing
Screened
interval
Land Surface
Water Table
Figure 4. Discrete multilevel sampler (DMLS) design and application in monitoring wells.
using discrete multi-level samplers (DMLS, Ronen
et al, 1986; Figure 4) or a specially designed discrete
interval sampler used in combination with a mini bladder
pump (Figure 5). Wells located outside of the PRB
were installed using either air-rotary or hollow stem
auger rigs. The majority of the wells installed in the
aquifer adjacent to the PRB are of similar construction
to the 5.08 cm ID wells within the PRB. All wells were
developed using pumping and surging techniques. Well
construction details for the performance monitoring
network are provided in Table Al (Appendix A). All
wells were surveyed into the existing site wide well
network using a Topcon Model CTS-2 Total Station
and location data for nearby wells were provided by
ASARCO.
Water Sampling and Analysis
A monitoring network of approximately 40 ground-water
sampling wells was installed in July 2005. Ground-
water samples were collected and analyzed at 1 month,
4 months, 12 months, 15 months, and 25 months of
operation. Ground-water samples were collected with
a submersible pump (Fultz Pumps, Inc.) or a mini
bladder pump (Innovative Sampling Systems). Flow
rates varied between 200 and 800 mL min'1 depending
on the well diameter and screen length. Samples
were collected following equilibration of geochemical
parameters: dissolved oxygen (DO), pH, oxidation-
reduction potential (ORP), and specific conductance
in a sealed flow-through cell (YSI 556). Measured
ORP values were converted to Eh values by adding the
difference between the measured ORP of a reference
solution (Orion ORP solution) and the theoretical ORP
of the reference solution. At the time of sampling,
turbidity values were generally less than 20 NTUs (mean
value 16 NTU, w=132) as determined using a Hach
turbidimeter (Model 21 OOP). Field measurements were
made for sulfide and ferrous iron using the methylene
blue and 1,10-phenanthroline colorimetric methods,
respectively (HachDR/2010). Alkalinity measurements
-------
Tubing from
bladder pump
Multi-electrode /
flow cell
Well casing —
Screened
interval
Land Surface
Water Table
Water in-let
head
Figure 5. Specially designed discrete interval sampler used with a mini bladder pump.
were made by titrating ground-water samples with
standardized 1.6 N H2SO4 to the bromcresol green-
methyl red endpoint.
Filtered samples (0.45 |am, Gelman Aquaprep) were
collected for metals and cation analysis and acidified to
pH<2 with ultra-pure HNO3. Analyte concentrations
were measured using inductively coupled plasma -
optical emission spectrometry (ICP-OES, Perkin-Elmer
Optima 3300DV). Most samples for arsenic speciation
were filtered, acidified with ultra-pure HC1, and retained
in amber-plastic bottles. Speciation analysis was carried
out using ion chromatography (1C) coupled on-line to
ICP-mass spectroscopy (IC-ICP-MS; Thermo Electron
Spectra HPLC). For samples in which thioarsenic
species were suspected based on elevated dissolved
sulfide concentrations (>0.2 mg I/1), filtered samples
were collected in amber glass bottles (precleaned
3000 class) and frozen (no acid added). Filtered and
unacidified samples were analyzed for major anions
by capillary electrophoresis (CE, Waters). Filtered
samples were also collected for dissolved organic carbon
(Dohrmann DC-80 Carbon Analyzer).
Field analyses were generally completed within
10 minutes of sample collection in order to minimize
any oxidation of dissolved ferrous iron and sulfide.
Electrodes used for geochemical parameters were
calibrated with certified buffer solutions and periodically
rechecked through daily sampling routines. Sample
bottles were kept refrigerated after collection and were
shipped back to the R.S. Kerr Environmental Research
Center (Ada, OK) in ice-packed coolers. Duplicate
samples were collected at a frequency of about 1 in
every 10 wells. Method reporting limits and results of
quality assurance/quality control (QA/QC) samples are
presented in Appendix B. In between wells, pump heads
and tubing were rinsed with distilled water. In selected
instances analysis of pump rinsate indicated generally
non-detectable concentrations of metals, cations, and
anions.
Arsenic Speciation Modeling
Equilibrium arsenic speciation modeling was carried
out using The Geochemist's Workbench, Release 6.0
(RockWare). Thermodynamic databases were modified
to include the As-O species in Nordstrom and Archer
(2003), As-S species presented in Wilkin et al. (2003),
As2S3 solubility products reported inEary (1992) and
Webster (1990), and ferrous arsenate phases reported
in Johnston and Singer (2007b). The standard database
(thermo.dat) was modified and used for modeling the
geochemical speciation of ground water (Appendix C).
-------
Core Sampling and Analysis
Core samples from the PRB were collected after
15 months to assess the uptake of arsenic and to evaluate
corrosion and mineral buildup on the iron surfaces. Core
collection methods and analysis procedures are described
in previous publications on PRB long-term performance
(U.S. EPA, 2003a,b). In all cases 5 cm inner diameter
cores were collected using a Geoprobe™. Core barrels
were driven using a pneumatic hammer to the desired
sampling location and continuous, up to 110 cm, sections
of iron or iron + soil were retrieved. Vertical cores were
collected in order to determine the spatial distribution
of mineral buildup in the reactive medium. Angle cores
were not collected; direct push methods were incapable
of advancing core barrels through the subsurface but
were amenable in the gravelly fill directly above the iron
and in the PRB. Core materials from the East Helena
PRB were black to gray in color without any obvious
signs of cementation or oxidation (Figure 6).
Figure 6. Photograph of a core segment collected
from the PRB after 15 months of operation
(September 2006).
Immediately after collection, the cores were frozen and
shipped back to the R.S. Kerr Environmental Research
Center for sub-sampling and analysis. The frozen cores
were partially thawed and then placed in an anaerobic
chamber with a maintained H2-N2 atmosphere. Each
core was logged and partitioned into 5 to 10 cm
segments. Each segment was homogenized by stirring
in the glove box and then split into 3 sub-samples for:
(1) inorganic carbon analyses, (2) sulfur analyses/x-ray
diffraction (XRD)/x-ray absorption spectroscopy, and
(3) Scanning electron microscopy (SEM). All sub-
samples were retained in airtight vials to prevent any air
oxidation of redox-sensitive constituents.
To determine elemental concentrations in bulk solids,
samples were digested in a microwave oven in 10%
nitric acid, and digestates were analyzed for metals
and non-metals by the same methods as those used for
ground water analysis. Concentrations of inorganic
carbon in core samples were determined with a carbon
coulometer system (UIC, Inc. Model CM5014).
Inorganic carbon analysis results are given in weight
percent C based upon carbon released from a sample
after acidification with hot 5% perchloric acid. This
acid digestion procedure releases inorganic carbon
present in minerals such as calcite (trigonal CaCO3),
aragonite (orthorhombic CaCO3), siderite (FeCO3),
magnesite (MgCO3), rhodochrosite (MnCO3), ferrous
carbonate hydroxide (Fe2(OH)2CO3), and carbonate
green rust (Fe6(OH)12CO3-2H2O). Measurements of total
sulfur in the solid phase were carried out using a sulfur
coulometer (UIC, Inc. Model CM5014S) in combustion
mode. Details of the sulfur measurement method are
described in Wilkin and Bischoff (2006).
Powder x-ray diffraction analysis of core samples
collected from the East Helena site was conducted
to determine the mineralogy of the aquifer materials
and precipitates formed in the iron treatment zones.
Materials for analysis were prepared by sonicating iron
core samples in acetone for 10 minutes followed by
filtration of the released particulates through 47 mm
diameter, 0.2-|am filter paper (polycarbonate). The
separated particles were mounted on a zero-background
quartz plate and scanned with Fe Ka radiation from 10°
to 90° 2-theta using a Rigaku Miniflex Diffractometer.
Scanning electron microscopy (SEM) and energy
dispersive x-ray spectroscopy (EDS) was used to
evaluate the morphology and composition of mineral
precipitates on the surfaces of zero-valent iron particles
collected at the East Helena site. Measurements
were conducted on polished samples to determine
the composition of surface precipitates on a semi-
quantitative basis. Samples for SEM and EDS
analyses were stored in an anaerobic glove box and
then embedded in an epoxy resin. The sample mounts
(2.5 cm diameter round mounts) were ground and
polished using diamond abrasives and coated with a
thin layer of gold prior to being placed within the SEM
sample chamber. Secondary electron and back-scattered
electron images were obtained using a JEOL JSM-6360
SEM. The instrument was operated using a 20 kV
electron accelerating potential and a beam current of
-------
about 10 nA. Micrographs were obtained at a range of
magnifications from about 50x to 5000x. Copper grids
obtained from SPI Supplies (West Chester, PA) were
used to verify quantitative length scales. EDS spectra
were acquired using an Oxford Instruments Model 6587
EDS Unit.
X-Ray Spectroscopy
X-ray absorption spectroscopy (XAS) measurements
were made at the Advanced Photon Source (Argonne
National Laboratory) in February 2005 and March
2007. The February 2005 XAS data were collected at
the PNC Collaborative Access Team (CAT), beam line
20-BM, Sector 20 and the March 2007 data at the DND-
CAT, Sector 5, beam line 5BM-D. Table 1 provides a
list of the samples/standards analyzed, the beam line
used to collect the spectroscopic data, and the type of
XAS data collected for each sample. The As K-edge
(11867 eV) measurements were collected using a silicon
(111) double-crystal monochromator at both beam
lines. The x-ray fluorescence signals were monitored
using a 13-element solid-state Ge detector at PNC-CAT
and a Canberra 13-element SSD detector at DND-CAT.
Transmission signals were collected at both beam lines
using ionization chambers. All XAS measurements
were performed at room temperature. Data for reference
samples were collected in transmission mode. Samples
were ground to a fine powder using an agate mortar
and pestle and evenly spread onto Kapton tape and
sealed with another piece of Kapton tape. Transmission
samples were inserted directly in the beam path between
the Io and ll ionization detectors prior to XAS data
collection. Samples collected in fluorescence mode
were packed into 1-mm thick plastic holders and sealed
with Kapton tape. The sample in the plastic holder
was placed in the beam path at a 45° angle to the
incident beam. The fluorescence detector to sample
distance was adjusted for each sample to maximize
the fluorescence signal at energies above the arsenic
absorption edge. Redox sensitive reference materials
and samples were handled in a N2 filled glove bag and
analysis was conducted in a plastic glove bag containing
a N2 atmosphere and continuously purged with N2 gas to
exclude O2.
As indicated in Table 1, the XAS analysis consisted of
x-ray absorption near edge spectroscopy (XANES).
Table 1 provides the data collection parameters for
the XANES analysis for data collected at both PNC-
CAT and DND-CAT. The energy calibration was
accomplished using metal foils or powdered reference
materials of known edge position. A gold foil (edge
position= 11918.7 eV) was used for data collected at
PNC-CAT. Data collected at DND-CAT used an arsenic
foil (edge position= 11866.7 eV) and sodium arsenate
(edge position= 11874.0 eV). The use of the arsenic
foil was abandoned because it was found that with time
the foil was oxidized by the x-ray beam and the edge
position shifted toward arsenite. Sodium arsenate is
an oxidized form of arsenic and does not exhibit beam
damage and was therefore stable and desirable as a
reference material for the arsenic absorption edge. It
also should be noted that the use of sodium arsenate
is advantageous over the gold foil because the edge is
closer to the absorption edge position of the unknown
samples.
Table 1. Samples analyzed using x-ray absorption
spectroscopy.
Sample Id
30ft
35ft
45ft
50ft
Core 1 30-34 ft fines
Core 1 44-48 ft
Core 2 30-34 ft
Core 2 3 8-42 ft fines
Core 2 3 8-45 ft fines
Core 3 30-38 ft
Core 3 38-42 ft
Core 3 3 8-42 ft fines
Core 4 30-34 ft
APBH4-2 10-11.5 ft
APBH 4-1 5-6.5 ft
TW 1-5 20-25 ft
TW 1-8 38-40 ft
TW 1-6 25-30 ft
Arsenopyrite
As2O3
Arsenian Pyrite
As(III) sorbed to
Ferrihydrite
As(V) sorbed to
Ferrihydrite
Elemental As
Sample Type
Boring
Boring
Boring
Boring
PRBCore
PRBCore
PRB Core
PRB Core
PRB Core
PRB Core
PRB Core
PRB Core
PRB Core
Source Zone
Source Zone
Source Zone
Source Zone
Source Zone
Reference
Reference
Reference
Reference
Reference
Reference
Beam
Line
PNC
PNC
PNC
PNC
DND
DND
DND
DND
DND
DND
DND
DND
DND
DND
DND
DND
DND
DND
DND
DND
DND
DND
DND
DND
XAS Data
XANES
XANES
XANES
XANES
XANES
XANES
XANES
XANES
XANES
XANES
XANES
XANES
XANES
XANES
XANES
XANES
XANES
XANES
XANES
XANES
XANES
XANES
XANES
XANES
As(III) sorbed FeS
As(III) sorbed FeS
Oxidized
As(III) sorbed
to FeS Oxidized
(duplicate)
As(III) sorbed to
Pyrite
Scorodite
Reference PNC
Reference DND XANES
Reference PNC XANES
Reference DND XANES
Reference DND XANES
Notes: PNC is Pacific Northwest Consortium; DND is
Dow-Northwestern-DuPont. These are beamline access
facilities at the Advanced Photo Source (Argonne National
Laboratory, Argonne, IL).
-------
Raw XANES data were processed using the Athena
software package (Ravel and Newville, 2005) that is
based on the IFEFFIT procedures (Newville, 2001).
Multiple scans (2 to 10 scans depending on data quality
and depending on whether the scans were collected in
fluorescence or transmission mode) of each sample were
collected and each scan was aligned using a reference
spectrum shown in Table 1. The aligned spectra were
merged into a final spectrum for each sample. The
merged averaged spectra were background corrected
and step height normalized. The normalized data were
used for XANES and linear combination of fits (LCF)
analysis. XANES and LCF operations were processed
using Athena. XANES analysis consisted of examining
the first derivative of the spectrum to determine the edge
position and comparing the edge position to standard
reference materials. The edge position of the standard
reference materials was used to help determine the local
bonding environment surrounding the arsenic atom
(see Table 2). Similarly, the white line position could
be determined using the raw spectrum and the white
line position could also be used to determine the local
bonding environment around the central arsenic atom
(Table 2).
Table 2. Absorption edge and white-line positions of
reference materials used in data analysis of the PRB
cores, source zone materials and well borings.
Compound
Elemental As
Orpiment
As(III) sorbed FeS
As(III) sorbed to
Ferrihydrite
Sodium Arsenite
Enargite
Sodium Arsenate
As(V) sorbed to
Ferrihydrite
Edge Position
(eV)
11866.7
11868.7
11867.6
11869.9
11870.1
11870.2
11874.0
11873.8
White Line
(eV)
11868.7
11870.1
11870.5
11871.9
11871.8
11871.7
11875.5
11875.5
LCF analysis was accomplished using the Athena
software package. The LCF analysis was used to
determine the types of bonding environments in
the samples. In addition, LCF can be used to give
the relative contribution of each bond type in the
analyzed samples. The LCF fitting was accomplished
using the normalized XANES over the fit range:
-20 eV < E0 < 30 eV The data were fitted using all
possible combinations of standards and the fit weights
were forced to be > 0. Initially LCF analysis was used to
determine the types of arsenic bonds present (e.g., As-S,
As-O) as well as the formal oxidation state of arsenic.
This type of fitting was accomplished using mineral
phases of known valence and bond type: elemental
arsenic (As(0)), sodium arsenite (As(III)-O), sodium
arsenate (As(V)-O), orpiment (As(III)-S), enargite
(4:1 As(V)-S), and dimethyl thioarsenate (2:1 As(V)-S,
DMTA). In all samples it was found that only As(V)-O,
As(III)-O, and As(III)-S had contributing fractions
above 0. Additional LCF analysis was carried out on
the samples using As(III) or As(V) sorbed to specific
iron mineral phases. The significant phases were
likely As(III) sorbed to ferrihydrite, As(V) sorbed to
ferrihydrite, and As(III) sorbed to FeS.
Hydrologic Methods
Hydraulic gradients at the water table were estimated
using both potentiometric surfaces interpreted from
manual ground-water elevation measurements and from
solution of the three-point problem using data from wells
instrumented with pressure transducers/data loggers.
A method similar to that described in Devlin (2003)
was used for automated solution of the three-point
problem. Gradients were estimated by fitting a plane to
the hydraulic head data using multiple linear regression
techniques. The hydraulic gradients calculated using the
three-point solution were compared with estimates of
gradients from potentiometric surfaces and found to be
similar (Table 3). This allowed gradients to be estimated
on a daily basis and averages more representative of the
full range of hydrologic conditions to be calculated.
Scorodite
11874.0
11875.8
Notes: Edge position is defined as the maximum in the first-
derivative of the energy vs. absorption function. White
line position is defined as the maximum in the energy vs.
absorption function.
-------
Table 3. Comparison of hydraulic gradients calculated
using multiple wells in the vicinity of the PRB and using
data from three wells (DH-17, PBTW-2, and EPA08).
Multi-Point Estimate Three-Point Estimate
Measurement
T-> t A , •. j Direction ,, . , Direction
Date Magnitude ,, . Magnitude ,,
(degrees) (degrees)
10/6/05
6/6/06
9/18/06
7/19/07
10/1/07
4/1/08
0.005
0.008
0.005
0.005
0.005
0.006
336
342
338
331
335
329
0.005
0.008
0.005
0.005
0.005
0.006
335
323
339
345
335
331
Field measurements of hydraulic conductivity were
made using single-well pumping tests and pneumatic
slug testing. Pneumatic slug tests within the PRB have
been performed using equipment similar to that depicted
in Figure 7. This method is based on recommendations
derived from Butler (1997) and utilizes air pressure and
vacuum to initiate instantaneous changes in head within
the well combined with high frequency monitoring of
the aquifer response using data loggers and pressure
transducers. This type of test provides the instantaneous
change in hydraulic head needed for the performance
of meaningful slug tests in media with high hydraulic
conductivity. The response data were analyzed using
the methods of Bouwer and Rice (1976), Springer and
Gelhar (1991), and Butler and Garnett (2000).
A sensitive electromagnetic borehole flowmeter was
used to define the relative hydraulic conductivity
Quick Release Valve
Land Surface U-il
Pneumatic Manifold Assembly
Regulator
Pressure / Vacuum
Pump
Wellhead Adapter
Water Table
Pressure Transducer
'Well Casing
Well Screen
Figure 7. Typical equipment used in performance of pneumatic slug tests.
-------
distribution of aquifer materials within and in the
vicinity of the pilot-PRB. The studies consisted of
measuring the vertical component of ground water
flow at fixed intervals in the wells under undisturbed
(ambient) and pumping conditions (Figure 8).
Measurements were made during constant-rate ground
water extraction to define the distribution of ground
water flow to the well. The rate of water flow to the
well from an individual interval is proportional to the
hydraulic conductivity of the materials adjacent to the
screen. Therefore, knowledge of the contribution of flow
from each measurement interval allows interpretation
of the hydraulic conductivity structure relative to the
average hydraulic conductivity of materials screened by
the well (Molz et al, 1994; Young et al., 1998).
Flowmeter
Alluvium
1 Pressure Transducer
Data Logger
Wall
Screen
LLJ
Flowrate
Volcanic Ash
Figure 8. Schematic diagram of electromagnetic
borehole flowmeter test design.
The electromagnetic borehole flowmeter used in
these studies was a commercially available system
manufactured by Tisco, Inc., consisting of a 1.3-cm
ID downhole probe, a 2.5-cm ID downhole probe,
and an electronics module. Probe design is based on
Faraday's Law which states that the voltage induced by
an electrical conductor moving through a magnetic field
is directly proportional to the velocity of the conductor.
The major components of the probe include an
electromagnet and a pair of electrodes mounted at right
angles to the poles of the magnet. The downhole probe
is designed as a hollow iron core through which water
flows. The electromagnet surrounding the core produces
a strong magnetic field. A voltage that is proportional to
the average water velocity is generated as the conductor
(i.e., ground water) flows through the magnetic field.
An uphole electronics package connected to the probe
amplifies and displays the voltage signal. Real-time data
acquisition is controlled by an associated computer. The
system is capable of measuring flow rates ranging from
less than 0.040 L miff1 to 40 L miff1.
Prior to conducting the study, the flowmeter was
calibrated in test cells constructed of materials identical
to those of well construction. Calibration was performed
by measuring water discharge from the test cell using
graduated cylinders and comparison with the associated
voltage measured by the flowmeter. Flow rates were
chosen to span the range of rates that would be used in
the field. Data obtained during the calibration phase
indicated that the meter responses for both the 1.3-cm
ID probe and the 2.5-cm ID probe were linear over the
range of potentially applicable flow rates.
The field tests were conducted using a procedure
based on the methods of Molz et al. (1994) and
Young et al. (1998). Flow rate measurements using
the electromagnetic borehole flowmeter were made
under ambient and constant-rate pumping conditions
at a measurement interval between 30 cm and 60 cm
within the well screen. The use of the 1.3-cm ID probe
was attempted for flow measurements under ambient
conditions. However, extreme variability, which was
likely due to electromagnetic interference from outside
sources, necessitated the use of the less sensitive,
2.5-cm ID probe which was used for all measurements
under pumping conditions. The tests were performed
at constant pumping rates ranging from 3 L miff1 to
7 L miff1. The rates were chosen to induce sufficient
flow from each test interval with negligible head loss
across the downhole probe. Each test was performed
using the following general protocol:
1. Ambient vertical flow rates were measured from
total depth to the top of the screen or water table
under static conditions.
2. The 2.5-cm ID probe was lowered to the bottom of
the well. A submersible pump was installed at the
water table and pumping was initiated to establish
a horizontal flow field. A pressure transducer and
data logger were used to monitor the water-level
response to groundwater extraction. The discharge
rate from the pump was measured using a graduated
cylinder and a stop watch at intervals not exceeding
approximately 30 min.
3. The establishment of a steady horizontal flow field
was indicated by stability in the pressure response to
-------
pumping which occurred after only a few minutes.
After conditions in the well stabilized, the flowmeter
was used to measure vertical flow rates at each of
the elevations occupied during the ambient flow
profile.
Measurements of vertical flow rates under ambient
and constant-rate pumping (induced flow) conditions
were analyzed using methods described by Molz
and Young (1993) and Young et al. (1998). The sign
convention used in this study is positive for upward
flow and negative for downward flow within the well.
The ambient flow rate at each measurement point is
subtracted from the flow rate measured at that elevation
under constant-rate pumping to obtain the portion of
total flow due only to pumping. The differences between
these pumping-induced flow rates at different elevations
represent the differences in horizontal flow toward the
well due to differences in hydraulic conductivity of
aquifer materials and hydraulic gradients. Assuming the
hydraulic head distribution along the well screen was
essentially uniform under the low flow rate conditions
of these tests, the relative hydraulic conductivity
distribution was then estimated using:
where: K is the average hydraulic conductivity of
screened materials, K. is the horizontal hydraulic
conductivity of interval /', Ag. is induced flow from
interval /', Aa is ambient flow from interval /', Az is the
' *i > i
thickness of interval /', QP is the total extraction rate,
and b is the aquifer thickness influenced by the test. The
estimated relative hydraulic conductivity distribution
may be converted to a distribution in units of volume per
time using an estimate of the average or bulk hydraulic
conductivity obtained from a pumping or slug test.
A natural gradient tracer test was also performed using
a dilute sodium bromide solution to provide a direct
line of evidence verifying ground-water flow through
the PRB. A slug of potable water amended with sodium
bromide was injected near the upgradient edge of the
PRB. The transport of the bromide pulse was monitored
through periodic sampling of ground water using wells
screened within the middle portion of the PRB between
elevations of approximately 1177.5 and 1178.0. Samples
were analyzed for bromide using Lachat flow injection
analysis.
QP/b
-------
6.0
Results and Discussion
Ground-water Geochemistry
A snapshot trend of major ion chemistry across the site,
from the primary source area for arsenic downgradient
to the PRB location and further to the site boundary,
is shown using a modified Durov diagram (Figure 9).
Ground water in the former speiss handling area
is slightly alkaline (pH ~10), and elevated in total
dissolved solids (TDS>3500 mg I/1) and arsenic
(>100 mg I/1). Sodium is the major cation and
approximately equimolar concentrations of bicarbonate
and sulfate are present in ground water near the source
area for arsenic. Downgradient from the former speiss
handling area, ground water evolves to lower pH (-6.5),
TDS values (<1500 mg I/1), and arsenic concentrations
(<50 mg I/1); sulfate becomes the dominant anion and
the proportion increases of calcium and magnesium
relative to sodium (Figure 9).
Long-term trends in the major ion chemistry of ground
water near the PRB test area are shown in a series of
Stiff diagrams on Figure 10 constructed using data from
well PBTW-1 (see Figure 3). These plots show little
variation over about six years in the proportions and
absolute concentrations of major cations and anions
sampled from monitoring well PBTW-1. In the area of
the pilot-PRB and over most of the site, ground water is
generally of Na+-SO42' type, withNa+ (150-900 mg I/1)
and SO/' (300-1000 mg I/1). On a molar basis,
sodium is enriched over calcium by about 6 times and
sulfate is enriched over bicarbonate by about 3 times.
Variations in ground water chemistry can be attributed to
interactions between ground water, aquifer minerals, and
wastes generated on the plant site.
Geochemical profiles across the entire saturated zone
for selected wells are shown on Figure 11. These depth
250T
1500 2000 2500 3000 3500
Total Dissolved Solids (mg/L)
Former speiss
handling area
Figure 9. Modified Durov diagram showing trends in major cations, anions, total dissolved solids, and arsenic
concentrations from the former speiss handling area to the northern site boundary (data collected June
2006).
-------
profiles were obtained using a multi-layer sampler with
baffles fit flush to the well casing, enabling sampling
of discrete depth intervals in the saturated zone (Ronen
et al, 1986). At well locations PBTW-1 and PBTW-2,
solute profiles are relatively uniform with no sharp
concentration gradients. Concentrations of total arsenic
vary between about 40 and 50 mg I/1 across the entire
saturated thickness in these wells. Arsenic speciation
results indicate that arsenite is the dominant form of
arsenic with some As(V) detected (2 to 14% of the total
arsenic, excluding the shallowest sampling point in
well PBTW-1). Wells DH-50 and EPA04 show more
pronounced chemical gradients (Figure llc,d). For
example, a 3-fold increase in arsenic concentration is
observed at the base of well DH-50, which corresponds
with a change in aquifer geology from a sandy matrix
to a sandy matrix mixed with coarse gravel and
cobbles. In well EPA04, located near the pilot-scale
PRB (Figure lie), a transition in arsenic concentration
is observed at a depth of about 12m below ground
surface, a concentration trend mirrored by sulfate and
specific conductance. Monitoring well results are also
shown on Figure 11 for samples collected at comparable
times for which the discrete interval sampling was
carried out. In most cases, the bulk well results give a
good representation of the average chemistry across the
saturated aquifer.
Arsenic concentrations in ground-water samples
collected over a period of 6 y vary from below detection
limits to a high value of 238 mg I/1. A comparison
between total dissolved arsenic concentrations and the
sum of arsenic species (arsenite + arsenate) is shown
on Figure 12. In over 90% of the samples, total arsenic
values are in very good agreement with the sum of
arsenic species (±10%). There appears to be a tendency
for the sum of arsenite plus arsenate to be slightly
less than the measured total arsenic value. This slight
discrepancy is likely related to a bias stemming from the
large dilutions often necessary to conduct the speciation
analysis on high concentration samples rather than to the
possibility of there being a missing species of arsenic
unaccounted for in the speciation analysis. As will be
discussed later, thioarsenic species are present in some
wells located within the PRB. Organic species of arsenic
were not detected during this study.
30
I
20
I
10
meq/kg
0
I
10
I
20
I
30
J
Cations
20
I
10
meq/kg
0
I
10
I
Anions
20
I
30
J
August 2001
Mn2*
r
rv,** J
30 20 10 C
I I I
August 2003
August 2002
10 20 30 30 20 10 C
III III
July 2005
10 20 30
I I 1
30
I
20
I
10
10
I
20
I
30
I
June 2006
Mn2*
7
x/
/
__/
30
I
20
I
10
10
20
I
July2007
SO,2" Mg*-
HCO, Ca"
a Na' + IC
30
so,2-
HCO,-
cr
Figure 10. Long-term trends in the major ion chemistry of ground water collected from monitoring well PBTW-1.
-------
a)
8.2-
Water
level 3
9.1-
_i
1 10.1-
co
O5
CO
1 11.0-
1
CO
1 11.9-
12.8-
13.7-
.
.
C.
\
7.3 -,
) Water
' level
8.8-
w
1 10-4-
E
q
" 11.9-
8
£
io 13.4-
•o
3
14.9-
16.5-
1
J
!
C
w
-- c
a if
Ne
r
-
ll
: a
i I
) t
^ K
ell PBTW-1
V
•*
X
.82 c
1 =
< i
HIPB
' . t "
V
'•
' •' i
'.*"
o
i
1
CY
S
d
10-
'
11-
E
Q 12-
13-
0
S
9
10
11
° 12
13
T)
D 14
0
J
5
TW-2
10-
11-
12-
^ 13-
14-
15-
0
S
10-,
11-
12-
13-
14-
f
) 16-
) C
j
o Purge value/Pumping depth
9
\ 10
\
f E"
4 |l2
\ -
V
• 1 • 1 • 1 • 1 • 1 • 1
pecific Conductance, mS/cm
9
1 10
f
. 11
\ E
\
i
0 500 1000 1500 2000
sulfate, mg/L
9
/\ '
\ \
I > > "
/ I
\° I"
V-
-^As / 13
1 -«-As3 j .
0 10 20 30 40 50
As, mg/L
-, 9
10
. I
. 1 i.
13
678
PH
o Purge value/Pumping depth
10 -|
" 11 -
• 12-
I
i S 14-1
\ 1
. 15-i
.
i 1
D 0.5 1.0 1.5 2.0 2.5 3.0 3.5 4.0 0
pecific Conductance, mS/cm
10 -,
I
. 12-
°} f13-
1 °,4-
j
•, 16-
10-,
II
1
1 M
^i Si4"
-•-AS J.r „.
•• I
10 20 30 40 50 0
As, mg/L
10-,
| 11-
1
A
,v a '
A 14-
1 1°"
y>
d •
)•
0 5 10 15 20 25
(Fe, Mn), mg/L
A
- i
\
A
0 \
1
00 150 200 250 300
Eh, mV
1 |-*-Mn
.
•
. J
, I
*
(Fe, Mn), mg/L
A
A
y
/*
1
500 1000 1500 2000 6.0 6.5 7.0 7.5 8.0 100 150 200 250 300
sulfate, mg/L pH Eh, mV
D J Water
' level
Figure 11. Geochemical profiles across the saturated aquifer in selected wells: a) PBTW-1, b) PBTW-2, c) EPA04, and
d) DH-50 (see Figure 3 for well locations).
-------
c)
d)
Well EPA04 ° Purge value/Pumping depth
WaterZ"
level
10.1-
i
to
< 11.0-
£
i
*~ 11 O-
T— II .9
I
3
" 12.8-
1
— '
13.7-
14.6-
"
-
~
-
-
-
•
-•
" 9~ •
• ' *•
, .
.••:•
;••••
.••:•
i-0.T
• •'•
•••• •
' A**
fW. .
•'•
fW. .
11 -
12-
£
° 13-
14-
1U-
t
1
*~^^^^
^~i~t 12-
\ ^
\ ° «-
\
T
/ ""
1 1 1 1 1 1 1 1 15
•
i 11
J
7 o 12
• i.
\ ° 13
1
'\
i 14
15
I— •— Fe "
*~f^^
~^.
0.0 0.2 0.4 0.6 0.8 1.0 01234 0-00 0.05 0.10 0.15
Specific Conductivity, mS/cm As, mg/L (Fe, Mn), mg/L
8.0 -|
8.5-
9.0-
° 9.5-
10.0-
£ D)
(0 •*•' *- 105
8.0
i'
8.5
;
i
• 9.0
I !"
^^^ 10.0
o \,.B
— 5 0 250 500 750 1000
-. 8.0-
-^_ 8.5-
~-~~~,M
I ^
* 9.5-
\ 0 10'°'
\
A
I
1
1
1
\
\ o
V
6 7 8 300 350 400 450 500
a) Q sulfate, mg/L pH Eh, mV
tils
(DO CT 5
O CO ^ Q
Figure 11. Geochemical profiles across the saturated aquifer in selected wells: a) PBTW-1, b) PBTW-2, c) EPA04, and
d) DH-50 (see Figure 3 for well locations).
-------
250
200-
150-
0-
n = 149
50 100 150 200
Total Dissolved Arsenic. mg/L
250
Figure 12. Comparison of total dissolved arsenic concentrations and the sum of arsenate plus arsenite.
Figure 13 shows a comparison between arsenic
concentrations in ground water and the pH-dependent
solubility of As2O3 and As2O5. This figure illustrates
that site ground water, even with high levels of arsenic
present, is still highly undersaturated with respect to
possible solubility controlling phases, As2O3 and As2O5.
In a following section this analysis is extended to
examine other possible phases that form in and around
the PRB, yet the trends indicate that factors controlling
arsenic concentrations in ground water are primarily
controlled by mineral-water reactions taking place in
the source area and dilution during transport, rather than
to retardation and arsenic attenuation by aquifer solids.
Redox conditions also impact the mobility and observed
arsenic concentration distributions. The highest total
arsenic concentrations are typically observed when
arsenic is dominated by arsenite (Figure 14); thus,
reducing conditions presumably play a role in governing
the plume dynamics. This correlation between reducing
conditions and elevated arsenic concentrations is also
shown on Figure 15 where ferrous iron concentrations
and Eh in ground water are contoured in the area around
the PRB. Note that ferrous iron and Eh values <150 mV
are typical in areas where arsenic concentrations are
high. Where ferrous iron concentrations are low
(<0.5 mg I/1), arsenic is generally present at lower
concentrations (<10 mg I/1) as arsenate.
™
o
ni
o
4-
2-
D-
-2-
-4-
-6-
•8-
-m-
XAs O. solubility
^
As.O solublty ^^^
Q) OO O
;^^^ 0 o o
° O O3S tK
0 o 0
- — 10ppbAs
10 11
pH
Figure 13. Arsenic concentrations in solutions saturated
with As2O5 and As2O3 and site ground water
(open red circles) compared to the MCL for
arsenic.
1.0-
= °6-
5
0.2-
Q»
ODO 0 CD 0 ,K) CD
o o
o
o
o o
10 20 30 «>
Total Dissolved Arsenic. mg/L
Figure 14. Comparison of total dissolved arsenic and the
fraction of total dissolved arsenic present as
arsenite, As(III).
-------
Code
•As >20 ppm
® As 10-20
®As 1-10
® No data
•
Figure 15. Distribution of redox indicators in ground water near the PRB: a) Fe(II) contoured from data collected in
8/2003, and b) Eh values in ground water in 7/2007.
PRB Behavior
Selected geochemical data for upgradient and in-wall
wells are presented in Table 4. Time trends of total
arsenic concentrations in monitoring wells located
hydraulically upgradient of the pilot-PRB are shown
in Figure 16. Well EPA02 is located on the eastern
upgradient edge of the PRB (Figure 3). Over the time
interval from 2002 to 2007, total arsenic concentrations
in well EPA02 have ranged between about 30 mg I/1
and 40 mg I/1. Most of the dissolved arsenic in this
well, 90-100%, is present in the reduced or trivalent
state. Well EPA08 was installed as part of the pilot-PRB
evaluation project after the PRB was constructed. This
well is located immediately upgradient and in the center
region of the pilot-PRB (see Figure 3). Sampling results
from well EPA08 show total arsenic concentrations
of approximately 28 mg I/1 entering the PRB. Again
arsenite is the major arsenic species in well EPA08
(>80%; Figure 16b).
In June 2006, a discrete multi-level sampler (DMLS)
was installed in upgradient well EPA08 for the purpose
of evaluating the depth-dependent concentration
trend in the aquifer. Results of this test are shown in
Figure 16c. Arsenic concentrations in well EPA08
vary little as a function of depth. The mean arsenic
value obtained from the DMLS study in well EPA08 is
26.5 mg I/1, which agrees reasonably with the bulk well
concentration noted above (-28.5 mg I/1; June 2006).
At the deepest sampling interval, ~1175 m AMSL, an
analytically significant increase in arsenic concentration
was observed, suggesting that in this area of the site
the highest arsenic concentrations in the aquifer are
present just above the basal ash tuff. A separate test
was performed in July 2007 using a mini-bladder pump
and baffled sampling assembly and similar results were
observed (Figure 16c).
Monitoring results for wells EPA02 and EPA08 show
that ground water entering the pilot-PRB is near-
neutral in pH (6.1-6.7), anoxic/suboxic (generally
<0.5 mg DO I/1), moderately reducing (Eh values from
130 to 200 mV), and contains moderate concentrations
of ferrous iron (<10 mg I/1) and dissolved organic
carbon (<5 mg I/1; see Table 4).
Ground-water sampling results from within the PRB
show evolving trends in the concentration of dissolved
organic carbon (DOC) and geochemical parameters that
clearly reflect interactions with zerovalent iron. For
example, Figure 17 shows time-dependent variations
in the concentrations of DOC, arsenic, pH and Eh in
ground water sampled from wells located within the
pilot-PRB compared to the influent. Notice that over
the first four months of operation DOC concentrations
exceeded 1,000 mg I/1. These high concentrations
-------
Table 4. Ground-water parameters in selected wells.
Parameter
Source
DH-33
DH-33
DH-21
DH-34
Upgradient
PBTW-1
EPA08
EPA08
EPA08, 33ft
EPA08, 37ft
EPA08, 42ft
PRB
T3A
T3A, early
T3A
T3A, late
S5
S8
S2
S4
Date
Jun-06
Jul-07
Jul-07
Jul-07
Jun-06
Jun-06
Sep-06
Jul-07
Jul-07
Jul-07
Jul-05
Jun-06
Sep-06
Jul-07
Jun-06
Jun-06
Jul-07
Jul-07
Al
(mg/L)
0.09
0.07
0.07
0.05
<0.05
0.23
0.25
1.30
0.28
0.22
0.03
0.07
<0.05
<0.05
0.146
0.113
<0.05
<0.05
Ca
(mg/L)
2.96
10.6
14.2
4.13
32.5
45.2
51.1
72.1
59.1
55.9
26.0
4.97
10.8
3.15
4.63
3.22
1.25
14.5
Mg
(mg/L)
0.836
8.43
6.82
1.78
9.45
18.3
20.3
28.4
23.4
22.5
7.94
1.55
12.6
3.73
0.61
4.95
<0.02
0.31
Mn
(mg/L)
0.10
0.31
0.62
0.02
1.4
4.7
5.1
9.4
6.2
5.8
0.99
0.04
0.07
0.01
0.03
0.01
0.00
0.01
Fe
(mg/L)
0.27
0.36
0.06
0.12
0.40
5.5
6.9
8.2
6.9
5.6
3.52
0.50
<0.01
<0.01
0.28
0.22
0.02
0.02
K
(mg/L)
6.41
5.85
14.5
19.0
11.9
13.9
13.1
15.1
14.4
14.5
3.4
9.5
10.5
15.1
4.59
7.3
11.1
17.6
Na
(mg/L)
786
869
807
548
509
249
220
223
222
228
126
255
281
233
95
120
194
155
Zn
(mg/L)
0.06
<0.02
0.03
<0.02
0.10
2.1
2.3
7.9
3.5
3.3
0.03
<0.02
<0.02
<0.02
0.03
<0.02
<0.02
<0.02
Si
(mg/L)
3.10
3.17
6.30
6.11
7.8
13.5
11.5
16.6
14.0
13.7
0.26
1.29
1.99
0.54
3.30
1.28
0.35
0.20
Cl
(mg/L)
60.0
74.8
25.4
30.2
67.3
42.5
37.2
34.0
33.5
34.8
38.8
50.3
42.6
34.6
23.0
19.7
23.8
17.5
so
(mg/L)
1050
1050
557
502
907
514
545
593
510
509
63
81
365
192
52
191
8.4
9.7
HC03
(mg/L)
672
917
1383
786
420
300
221
180
196
206
325
385
372
366
175
106
368
337
T
(°C)
11.7
11.7
10.9
11.1
13.4
13.6
13.1
13.2
13.2
13.2
nm
13.9
13.2
16.6
13.5
14.2
15.9
14.0
PH
(SU)
10.2
9.39
8.81
9.58
7.50
6.50
6.20
5.85
5.89
6.25
7.42
10.10
9.34
10.07
9.10
6.80
10.31
9.92
Eh
(mV)
157
15
189
68
173
207
159
210
221
224
-382
42
-43
181
113
25
-134
-75
DO
(mg/L)
0.3
0.5
3.6
0.6
0.2
0.1
<0.1
1.0
0.5
0.4
<0.1
<0.1
<0.1
0.3
<0.1
<0.1
0.4
0.5
SC
3.03
3.29
2.73
2.38
2.04
1.33
1.48
1.58
1.50
1.49
1.25
1.03
1.44
0.99
0.45
0.63
0.93
0.75
TOO
(mg/L)
7.14
9.45
21.4
5.88
4.50
3.62
3.11
6.12
3.58
3.96
905
20.4
4.38
7.17
9.69
4.98
28.0
171
As
(mg/L)
103
104
89
99
45
28.5
27.3
24.5
25.1
25.6
<0.1
0.07
0.25
0.15
0.15
0.11
0.03
0.02
As(lll)
(mg/L)
87
105
47
101
40.4
20.2
19.3
22.2
25.2
25.5
<0.01
0.03
0.11
0.19
0.08
0.09
<0.01
<0.01
As(V)
(mg/L)
1.26
<0.01
47
<0.01
0.95
4.8
5.6
0.36
0.19
0.12
<0.01
<0.01
0.07
<0.01
0.06
0.02
<0.01
<0.01
-------
"3> 4^ '
it
o 30
| 20-
1
B
J
EPAOZ O A
" °-o .
^~o
EPAOe
t
PRB
Installed
a)
1 13"*
1181-
1180-
j 1179-
V>
3/29/2002 3O6/2003 3^6/2004 3'26/20G5 3/26COO8 3/28/2007 ^ 1178-
1.0-
0.8-
~m 0-6-
c
o
= 0.4-
e
111
0,2-
00-
_^--*— — — -^ f*
*^~ EPA02 ~~ «)a|^ ^/
^__^-O
b)
_o
I 1177.
UJ
1175-
11 74
2
Depth-dependent
As concentrations in
well EPA08
Results .
from 7/07 .
"~~~^^ --^ Results
^*~°*" « from 6/06
• s
;
yi
1
'^
'*- _ t
c)
B 25 30 3
Total Dissolved Arsenic, mg/L
Figure 16. Arsenic concentration trends in ground water upgradient from the PRB: a) total dissolved arsenic as a
function of time in ground water from monitoring wells EPA02 and EPA08, b) fraction of total dissolved
arsenic as arsenite, and c) depth-resolved concentration trends in monitoring well EPA08.
*L.
1000.
j aoo.
~ 600-
{
3 400-
0.
(
r^
°^
10.
I
a.
9-
Upgradient concentrations <5 mg/L
ft °
e o
n=24 H=I9
60 8
2 4 6 8 10 12 14 16 18 20 22 24 26 28 3
Months of Operation
O
8
88 °
o
Q H 0
e § S|
^\^9^*>^>^
b)
f 3
I2.
£
<
0-
0 (
d)
'.00.
0-
> .100.
£.200.
-300.
•400.
Upgradient concentrations >25 mg/L
'
n»19 n-25
O O
11=24
0 ° §
Q O
9 i H h •-
2 4 6 8 10 12 14 16 IB 20 22 24 26 28 31
Months of Operation
'/^A^/^^/A^^////////,
° H
°
• ^ !
O o
« §
e
Months of Operation
Months of Operation
Figure 17. Plots showing trends with time: a) DOC, b) arsenic concentrations, c) pH, and d) Eh for wells sampling
ground water from the pilot-PRB.
-------
of DOC were likely residual carbon leftover from
the guar bio-slurry. During the first two sampling
events, ground water sampled from within the pilot-
PRB was difficult to filter (using 0.45 |am disc filters)
and slightly malodorous. After 12 months, however,
DOC concentrations decreased to near-background
levels (mean value 9 mg I/1), and after 15 months
DOC concentrations in the PRB were indistinguishable
from concentrations present upgradient of the PRB
(mean value 5 mgl/1). Consequently, approximately
1 year passed before the guar had fully broken down or
passed through the PRB. Generally low values of DOC
were also observed after 25 months, yet in some wells
increases in DOC concentrations were observed, perhaps
related to changing water levels or subsurface movement
of residual guar gum.
Concentrations of arsenic entering the PRB are
>25 mg I/1. Within the PRB, arsenic concentrations
have generally been <0.50 mg I/1. Of 80 samples
collected from the pilot-PRB, only 11 samples exceeded
0.50 mg As I/1 (see Figure 17b); 62 samples had
concentrations of arsenic at or below 0.50 mg I/1; and,
24 samples were at or below the MCL for arsenic of
0.01 mg I/1. The highest concentrations of arsenic
appear to be restricted to the deeper wells within the
PRB and this is likely reflective of regions where
ground-water seepage velocities are fast and/or where
less iron was emplaced in the subsurface possibly due to
collapse of the trench walls. Interestingly, the oxidation
state of arsenic within the PRB varies widely. The
fractional abundance of As(III) ranges from 0.20 to
1.00, with a mean value of 0.52, which is low compared
to the mean oxidation state of arsenic entering the
PRB (fractional abundance of arsenite >0.8). This
observation is consistent with laboratory tests with
arsenite and zerovalent iron that often show some
degree of arsenite oxidation to arsenate, perhaps due to
oxidation reactions with iron corrosion products (e.g.,
Manning et al., 2002; Lien and Wilkin, 2005). Overall
the arsenic concentration distribution inside the PRB is
similar between the 15 and 25 month sampling periods.
Geochemical parameters, pH and Eh, in the PRB show
expected trends. For example, the pH of ground-water
entering the PRB ranged from about 6.1 to 6.5. Whereas
the pH of ground water samples collected from the PRB
ranged from 6.2 to 10.8 (Figure 17c). As water reacts
with zerovalent iron the pH is expected to increase due
to the reaction: Fe(0) + 2H2O = Fe2+ + 2OH~ + H2. The
Eh of ground water entering the PRB ranges from about
130 to 200 mV, that is, the ground water is moderately
reducing. Within the PRB, Eh values ranged from about
140 mV to values as low as -380 mV. The lowest Eh
values (highly reducing) were recorded within the first
4 months of operation. After about 1 y, and continuing
through the second year of operation, Eh values
increased to a range typical for zerovalent iron PRB
systems, 100 to -200 mV (U.S. EPA, 2003a). Regions
of limited pH and Eh changes can be attributed to high
velocity and short residence time in the PRB, whereas
increases in pH and decreases in Eh are indicative of
longer reaction times between ground water and the
reactive medium (e.g., Liang et al., 2005).
Figure 18 shows an Eh-pH diagram for arsenic with
points plotted that contrast ground water collected from
the plume and the PRB. The data points also indicate
the measured speciation of arsenic in order to compare
analytical results with an equilibrium model. Within
the plume, pH is near-neutral and measured Eh values
would suggest the presence of mainly H2AsO4", HAsO42",
and H3AsO3°. This expected distribution of As(V) and
As(III) is generally observed, in that samples dominated
by arsenate tend to be elevated in Eh, whereas arsenite
is dominant in low Eh samples. Note in many cases,
however, measured arsenic speciation was dominated
by arsenite, but the equilibrium model, based on the
Pt-electrode response, predicted arsenate dominance.
Similar trends are observed for ground water collected
from the PRB, although higher pH values are present
in many cases. Interestingly, several measured points
in the PRB (after 1 month of operation) indicate highly
reducing conditions and fall within the stability field for
elemental arsenic. These highly reducing conditions
were not sustained past the very earliest operation of the
PRB.
>
iS
-.5
PRB
Plume
O ASM Ml) mixture • >80%As(V)
>80%As(IM) OAs(V, III) mixture
• >80%As(lll)
PO,=1 bar
0 2 4 6 8 10 12 14
PH
Figure 18. Eh-pH diagram for arsenic. Data points
show measured pH and Eh for ground water
in the PRB and plume; color code of points
shows the measured speciation of arsenic.
Diagram constructed for 25°C and SAs=10~4.
-------
Oxidation-reduction potentials were calculated from the
measured arsenic speciation based on the following half-
cell reaction:
H3AsO4°
2H+
2e- = H3AsO3°
H2O.
The activities of H3AsO4° and H3AsO3° were calculated
based on the measured arsenic speciation, pH, and
calculated ionic strength. The equilibrium expression
was solved for a- and the Eh was calculated using
the relation Eh=0.059pe. Because of the high
concentrations of arsenic in ground water at this site, this
system perhaps represents a best case scenario for the
Pt-electrode response to reflect aqueous distributions of
arsenate and arsenite. The calculated redox potentials
are compared with measured potentials in Figure 19. A
correlation exists, but there is poor agreement between
the speciation measured using IC-ICP-MS and speciation
predicted using the Pt-electrode readings. In all cases,
calculated redox values based on analysis of aqueous
speciation are lower when compared to Eh measurements
made with a Pt electrode. Results on Figure 19 are
also shown for studies by Holm and Curtiss (1989)
and Rude and Wohnlich (2000). Consistent with these
previous studies, results here indicate that Pt-electrode
response is not highly indicative of arsenic speciation in
ground-water samples, not even in this seemingly ideal
environment where elevated concentrations of arsenate
and arsenite might be expected to poise the oxidation-
reduction potential.
300
ffi"
200-
100-
0-
-100-
-200-
-300-
-400-
-500-
Rude and Wohnlich (2000)
Holm and Curttes (1989) Q
i9 *• •
§o*
Eh in aquifer
Eh in PRB
-200 -100 0 100 200 300 400 500 600 700
Eh, measured (mV)
Figure 19. Comparison of Eh values of ground water
measured using a platinum electrode and
calculated using the As(III)-As(V) redox
couple.
in June 2006, September 2006, and July 2007 at 12,
15, and 25 months of operation, respectively. The
concentration of total dissolved arsenic at these sampling
times was 27.3 mg L"1, 22.6 mg I/1, and 15.5 mg I/1,
respectively. Considering the low concentrations
of arsenic determined within the PRB from the first
sampling event at 1 month, the elevated concentrations
of arsenic in EPA09 determined in June 2006 and
July 2007 were unexpected. A more significant decrease
in arsenic concentration downgradient of the pilot-PRB
was expected. Consequently, detailed, depth-resolved
sampling was carried out in September 2006 and
July 2007. Note that the bulk well concentration in
EPA09 in September 2006 was 83% of the arsenic
concentration entering the PRB at that time. This value
decreased to 61% in July 2007. Results of the depth-
resolved sampling are shown in Figure 20. Arsenic
concentrations on the downgradient side of the pilot-PRB
are significantly reduced from the water table down to a
depth of about 1177.1 m AMSL. Arsenic concentrations
in this shallow aquifer interval were 0.07 to 0.65 mg L"1,
and generally coincide with concentrations observed
within the reactive medium (see Figure 17). Samples
taken at 1176.2 and 1175.3 m AMSL showed arsenic
concentrations of 20 to 30 mg L'1, respectively. These
results indicate that ground water is moving beneath
the pilot-PRB and transporting arsenic across the plane
of the PRB. In the upper region of the aquifer where
ground water is moving through the PRB, -99% arsenic
removal is achieved.
Hydraulic Investigation
The elevation of the water table in the vicinity of the
PRB varied within a range of approximately 1 m to
2 m on a seasonal basis (Figure 21 and Appendix D)
between June 2002 and April 2008. This fluctuation
likely reflects seasonal variations in infiltration from
Prickly Pear Creek and associated Upper Lake as well
as infiltration of precipitation. Since installation of
the PRB, ground-water elevations measured in well
TR8 (see Figure 3) within the PRB have ranged from
approximately 1179 m AMSL to 1181 m AMSL. This
indicates that the water table has remained below the
top of the zero-valent iron which is at an approximate
elevation of 1183.5 m AMSL.
Monitoring well EPA09 is located immediately
downgradient of the pilot-PRB. This well was sampled
-------
1182
Arsenic, mg/L
0 5 10 15 20 25 30 35
CO
5 1180-
E
o
> 1178-
_£
UJ
1176-
EPA08
Water level
Iron
Weathered ash
Arsenic, mg/L
0 5 10 15 20 25 30 35
1182-1—i—i—•—i—•—i—i—i—•—i—•—i—i-
1180-
117E
1176-
EPA09 * •
Figure 20. Depth dependent arsenic concentrations in wells EPA08 and EPA09 and conceptual model of ground-water
transport underneath the pilot-PRB. (Blue filled circles in EPA09 are from September 2006; red open
circles are from July 2007).
1181.50 -
CO 1161-00
§ 118050 -
LU
0) 1180.00 •
2 1179,50
1179.00 •
\ \ \ \
2002 2003 2004 2005 2006 2007 2008
Date
Figure 21. Ground-water elevations obtained using data from a pressure transducer/data logger installed in well DH-
17 located approximately 30 m upgradient of the PRB.
-------
Temporal variations in the potentiometric surface
near the PRB were interpreted from data obtained by
manual measurements of ground-water elevations on
nine dates between July 2005 and June 2008. Based
on these data, there is no evidence of ground-water
mounding upgradient of the PRB that would indicate
the PRB is acting as a significant impediment to ground-
water flow. The surfaces (Appendix E) were similar
to that interpreted from data obtained on April 1, 2008
(Figure 22), indicating that the hydraulic gradient was
relatively stable in both magnitude and direction through
time. In each case, the data indicated a northwest
direction of flow. The magnitude of the hydraulic
gradient near the PRB ranged from approximately 0.005
to 0.008 with an average of 0.006.
262500
,§
O)
262450
262400
414350
414400
414450
Easting (m)
Figure 22. Shallow potentiometric surface interpreted from ground-water elevation measurements obtained using an
electronic water level indicator on April 1, 2008. The contour interval between equipotential lines is 0.1 m.
The approximate location of the PRB is depicted in red.
To better evaluate the full range of variation in
hydraulic gradient potentially associated with seasonal
changes in ground-water elevations, data obtained from
pressure transducers/data loggers installed in wells
DH-17, PBTW-2, and EPA06 (Figure 23) were used
to estimate the direction and magnitude of the gradient
approximately every four hours between July 2005 and
April 2008. The estimated magnitude and direction
of the gradient were averaged on a daily basis. The
magnitude of the gradient calculated from this expanded
data set also ranged from approximately 0.005 to 0.008
(Figure 24) with an average of 0.006. The estimated
direction of ground-water flow ranged from an azimuth
of 322 deg to 350 deg (Figure 23) with an average
azimuth of 334 deg.
-------
262500-
3- 2624504
0)
262400
262350
414350
414400
414450
414500
Easting (m)
Figure 23. Wells instrumented with pressure transducers/data loggers and used to characterize hydraulic gradient
fluctuations on a daily basis. The azimuthal distribution in degrees of the calculated hydraulic gradient
vectors is depicted using a Rose diagram in the figure inset. Locations of the wells and approximate
location of the PRB (red line) are overlain on the potentiometric surface on April 1, 2008, for reference.
The contour interval between equipotential lines is 0.1 m.
A natural gradient tracer test was performed from
June 23 to 29, 2008, using a solution of potable water
amended with sodium bromide to provide another direct
line of evidence verifying ground-water flow through
the PRB. A slug of approximately 95 L of potable water
with a bromide ion concentration of approximately
500 mg L"1 was injected into well TR1 (Figure 25)
over a period of approximately 5 min. The injection
was restricted to a 0.9 m zone between approximately
1177.4 m and 1178.3 m AMSL by isolating this zone
with mechanical packers. The injection pipe consisted
of hand perforated 3/4" Schedule 40 PVC.
300 -,
•5 100 —
0.004
0.005 0.006 0-007 0.008
Hydraulic Gradient Magnitude
0009
Figure 24. Distribution of the magnitude of the
hydraulic gradient near the PRB estimated
on a daily basis using data obtained from
pressure transducers/data loggers installed in
wells DH-17, PBTW-2, and EPA06.
-------
The downgradient transport of the bromide was
monitored using nine observation wells (TR3, TR4-2,
TR5-2, TR6-2, TR7, TR8, TRIO, TR11-2, and TR12-2)
located in the PRB (Figure 25). Each of the observation
wells was screened in approximately the same interval
as the injection zone. Ground-water samples were
obtained at frequent intervals and analyzed for bromide
(Figure 26) using a Lachat flow injection analyzer.
3 m
TR11-2
TR10 \ TR12-2
TR9
EPA01
TRO TR1TR2
T1A
T1B
• 1-inch well
O 2-inch well © Multiport well
Figure 25. Ground-water monitoring wells within the
tracer test area of the PRB. The well used for
tracer injection, TR1, is highlighted in red.
o
0.0 —f""i " |—«-i—*|- - i— j —i r-
12345
Time Since Injection (d)
Figure 26. Ratio of the measured bromide ion
concentration to the concentration in the
injected slug as a function of time.
With the exception of well TR8, the bromide tracer
was observed at all locations. The first arrival of the
bromide pulse required approximately three days to
migrate across the PRB. A plot of the time required for
first arrival of the tracer as a function of distance from
the injection well (Figure 27) indicates the presence
of a heterogeneous flow field within the PRB with
more rapid transport to wells TRIO and TR11-2 and
slower transport to well TR6-2 than to the majority of
the wells in the network. Assuming that bromide is
conservatively transported in this system, an estimate
of the ground-water velocity through relatively fast
pathways based on a simple linear regression of the first
arrival times at wells TR3, TR4-2, TR5-2, TR7, and
TR12 is approximately 0.35 m d"1. These results confirm
that ground water moves at a significant rate through the
PRB.
1.5-1
1.0-
g
I
m
en
Q
0.5-
0.0
TR10
+
TR12-2
TKB-2
= 0.35*X + 0.45
R2=0.99
1 2
Time of First Bromide Arrival (d)
Figure 27. Distance of tracer migration as a function of
the time required for the first arrival at each
monitoring location. A linear regression was
performed using data from wells TR3, TR4-2,
TR5-2, TR7, and TR12-2.
A short-term, single well pumping test was performed at
well PBTW-2 to estimate bulk hydraulic conductivity of
the aquifer materials in this area during the PRB design
phase. Estimates of hydraulic conductivity ranged from
25.3 md"1 to 41.1 md"1 with an average of 33.2 md"1.
This estimate is supported by the value of 36.6 m d"1
obtained from pneumatic slug tests at this well. For
purposes of this investigation, a value of 36.6 m d"1 for
the bulk hydraulic conductivity of aquifer materials
adjacent to the PRB was used to estimate the interval-
specific hydraulic conductivity of these materials from
the borehole flowmeter surveys.
-------
Results of pneumatic slug tests performed in short
(0.76 m) screened wells located within the PRB (Table 5,
Figure 28) ranged from 2.1 m d"1 to 110 m d"1 with
an average of approximately 58 m d"1. The highest
estimates for the hydraulic conductivity of the PRB
were consistently obtained from wells screened below
approximately 1177.4 m AMSL. No changes were
observed in the distribution of hydraulic conductivity
between measurements made after 5 months and after
16 months (Figure 29). Pneumatic slug tests were
also performed in July 2007 in two wells, TR1 and
EPA10, that screen the majority of the PRB and are
located (see Figure 3) near the western end and center
of the PRB, respectively. Results indicate the average
hydraulic conductivity of materials adjacent to these
wells is approximately 43 m d"1 (well TR1) and 49 m d"1
(well EPA10). It is noted that the majority of these
values are equal to or significantly greater than the
permeability estimated for a sample of the zerovalent
iron material (30 m d"1 to 58 m d"1) using laboratory
test method ASTM 2434 during the PRB design phase.
This indicates that the PRB installation method did not
significantly reduce the hydraulic conductivity of the
PRB materials.
The borehole flowmeter methodology has been used
to characterize the hydraulic conductivity of aquifer
materials adjacent to six wells (PBTW-1, PBTW-2,
DH-17, EPA02, EPA08, EPA09, and EPA10) (Figure 3).
With the exception of existing well DH-17, each of these
wells is screened from approximately the water table
to the top of the ash unit underlying the upper aquifer.
Three of the wells bound the study area to the east
(PBTW-1 and PBTW-2) and south (DH-17), providing
information concerning the scale of heterogeneity in
this portion of the site. At well PBTW-1 (Figure 30),
aquifer materials immediately above the ash unit at
the bottom of the well screen appear to have higher
hydraulic conductivity than the average for materials in
this portion of the site.
Table 5. Estimates of hydraulic conductivity in units of
m d"1 obtained from pneumatic slug tests
performed in wells located within the PRB.
Well
EPA10
SOI
S02
S03
S04
S05
S06
S07
SOS
TR1
TR2
TR3
TR7
TR8
TR9
TRIO
October
2005
98
7.3
56
100
95
70
51
September
2006
88
75
98
11
42
110
92
69
32
55
43
47
2.1
34
July 2007
49
43
-------
C/3
1182 -
1181 -
1180
1179 —
1178
"S 1177 •
c
o>
0)
O 1176
GO
1175
1174
S04
Average Water Table Elevation
TR2 S02
TR10 TR7 TR8 TR3 308
807 S06
SO I
Approximate Bottom of Zero-Valent Iron
Range of Laboratory
Permeameter Results
40 80
Hydraulic Conductivity (mid)
120
Figure 28. Results of pneumatic slug tests within the PRB using a series of wells each screening a 0.76 m vertical
interval. The screened elevations are depicted by the length of the line representing each well. The results
of laboratory permeameter measurements of the zero-valent iron used in construction of the PRB are
provided for reference.
>» 2-
O ^
c
o
16 months
n=7
150
200
50 100 150
Hydraulic Conductivity, m/d
200
Figure 29. Results of pneumatic slug tests within the PRB after 5 months and 16 months.
-------
1181
1180
O 117
Top' of Ash
K (m/d)
K(m/d)
K (m/d)
Figure 30. Hydraulic conductivity distribution
estimated for materials adjacent to the
screened intervals of wells DH-17, PBTW-1,
and PBTW-2 based on characterization using
borehole flowmeter techniques.
During this test, consistently anomalous readings were
observed at approximately 1177.0 m and 1178.8 m
AMSL. The source of these anomalies cannot be
determined but may be related to well construction
problems such as void space between the well screen and
borehole wall in these zones. Due to the significance of
these anomalies, the values for hydraulic conductivity
of intervals above 1177.0 m AMSL may not be accurate.
At well PBTW-2 (Figure 30), aquifer materials between
depths of approximately 1174.5 m and 1176.4 m
AMSL appear to be between two and three times more
conductive than the average for all materials adjacent to
the well.
At well DH-17 (Figure 30), which bounds the southern
portion of the study area, aquifer materials in the lower
half of the screened zone appear to be as much as
three times more conductive than the average for all
materials adjacent to the well. This pattern is similar
to that observed in the other test wells. However, the
partially penetrating construction of this well may result
in increased flow in the measurements obtained near the
bottom of the well screen. Therefore, absolute values
of hydraulic conductivity for aquifer materials at the
bottom of the well screen may be overestimated.
The borehole flowmeter was also used to characterize
hydraulic conductivity in wells EPA02, EPA08, EPA09,
and EPA10 within and adjacent to the PRB (Figure 31).
At well EPA02, the most permeable materials appear
to be located at the top of the ash unit. Materials near
the water table were also more conductive than the bulk
hydraulic conductivity. However, anomalous readings
were observed at approximately 1176.0 m and 1177.3 m
AMSL. Therefore, the hydraulic conductivity estimated
for this interval represents an average value for materials
within this zone. The results from wells EPA08 and
EPA09 were similar with materials having hydraulic
conductivities significantly greater than the bulk
conductivity for the aquifer adjacent to the bottom half
of the PRB. Materials near the water table had hydraulic
conductivities lower than the bulk conductivity for the
aquifer. Although heterogeneity is evident, the hydraulic
conductivity profiles obtained from these wells display
similar patterns with respect to the existence of materials
with high hydraulic conductivity in the deeper portion of
the aquifer.
EPA 08
1 I
B Bottom I
• cfPRB •
0 '30 20C C 100 200 0 1CO 2CO 0 '00 20CI
K (m/d) K(m/d) K (m/d) K (m/d)
Figure 31. Hydraulic conductivity distribution estimated
for materials adjacent to the screened
intervals of wells EPA02, EPA08, EPA09,
and EPA 10 based on characterization using
borehole flowmeter techniques.
The borehole flowmeter technique was also applied to
characterization of materials adjacent to well EPA 10
(Figure 31) located near the center of the PRB and
screened across the entire saturated thickness of the
PRB. At this location, the hydraulic conductivity of
the PRB appears to increase with depth displaying a
trend that is generally similar to the results of estimates
obtained using pneumatic slug tests (Figure 28). These
data indicate that the hydraulic conductivity of the PRB
is generally equal to or greater than that of the aquifer
in most intervals. Although the results of the flowmeter
analysis at EPA 10 are considered to be somewhat
uncertain due to possible effects related to the proximity
of the trench walls, a comparison of the slug test data
with the estimated hydraulic conductivity profile for
well EPA08 (Figure 32) supports the conclusion that
the PRB does not provide a significant impediment to
groundwater flow.
-------
1181
1180
EPA08 Flowmeter Results
Slug Tests within PRB
CO
1179
1178
I 1177
OJ
ffl
LU
1176
1175
1174
T
40 80
Hydraulic Conductivity (m/d)
1
120
Figure 32. Comparison of hydraulic conductivity distribution adjacent to well EPA08 estimated using borehole
flowmeter techniques and the results of slug tests performed in a series short-screened wells within the
PRB. The screened intervals are depicted by the length of the line representing each well.
In summary, the hydraulic conductivity of the materials
adjacent to these wells, including the location within the
PRB, appears to be non-uniform. Materials in the lower
portion of the aquifer and PRB appear to be significantly
more transmissive than materials in the upper portion
of the saturated zone. This implies that ground-water
velocities and associated contaminant fluxes may be
significantly higher in the lower portion of the PRB.
Flux Evaluations
Data collected in the geochemical and hydrologic aspects
of this study can be combined to determine arsenic flux
in the subsurface. The mass flux of arsenic entering and
leaving the PRB was estimated by multiplying the depth-
dependent volumetric flow velocity (Darcy velocity) by
depth-dependent arsenic concentrations in wells EPA08
and EPA09 (Figure 33):
Arsenic Flux = (K • dhldx) C,
where K is the interval specific hydraulic conductivity,
dhldx is the hydraulic gradient, and C is the total
arsenic concentration at specific depths in ground water.
Volumetric flows were determined at the midpoint of
depth intervals tested with the borehole flowmeter.
Arsenic concentrations at these depth intervals were
extrapolated from discrete interval sampling profiles.
Arsenic mass flux entering the PRB ranges from 0.5
to 16.9 g nrM"1 with a maximum value at a depth of
about 1177 m AMSL. The flux of arsenic downgradient
of the PRB ranges from 0.01 to 20 g m"2d"'. Arsenic
mass flux is significantly reduced downgradient of
the PRB from the ground-water table to a depth of
about 1177 m AMSL. Deeper in the aquifer, where
emplacement of zerovalent iron was incomplete, there
is no indication that arsenic transport in the subsurface
is being impacted. Results plotted on Figure 33 clearly
show the impact of the PRB on arsenic attenuation.
The depth related analysis shows that specific intervals
with high contaminant flux need to be incorporated
into PRB design. Regions of high contaminant flux
are likely to determine the overall efficiency of PRB
systems for reducing contaminant transport. An
important implication is that bulk characterization of
wells screened across the saturated thickness may not
be able to resolve variability of contaminant flux in
the subsurface leading to inappropriate system design.
Clearly a full-scale PRB at the East Helena site requires
emplacement of zerovalent down to the confining ash
tuff layer with consideration of the high flux zone at
1177m AMSL.
-------
1180-
| 1179'
E 1178-
c
.2
CD
£ 1177-
1175
Arsenic flux in, EPA08
Arsenic flux out, EPA09
5 10 15 20
Arsenic Flux, g/m" d
25
Figure 33. Estimation of arsenic flux entering and
leaving the PRB as a function of depth.
PRB Performance
A number of factors have been shown to be important in
controlling the long-term performance of zerovalent iron
reactive barriers. Key among these factors are reaction
processes that lead to excessive oxidative corrosion
of iron granules and/or excessive carbonate mineral
precipitation (e.g., U.S. EPA, 2003a, Liang et al., 2005;
Li et al., 2006). Both of these processes degrade the
reactive and hydraulic performance of zerovalent iron
PRBs. Dissolved oxygen (DO), for example, rapidly
reacts with zerovalent iron and drives the formation
of various ferric oxide, oxyhydroxide, and hydroxide
minerals. In situations where ground water containing
several mg L'1 of DO interacts with zerovalent iron,
rapid cementation and loss of pore space and hydraulic
conductivity occurs due to the precipitation of ferric-
iron bearing minerals. Although the formation of these
Fe(III) corrosion products may be a benefit from the
standpoint of arsenic removal (e.g., Furukawa et al.,
2002), excessive corrosion can lead to the complete
bypass of the reactive medium. This particular problem
is not a concern at the East Helena Site. As noted above
DO concentrations entering the PRB are very low and
core samples show no evidence of nodule formation or
cementation after 15 months of operation.
Carbonate mineral precipitation in PRBs is primarily
driven by the pH increase that comes about during
anaerobic corrosion of metallic iron. Typically ground
water is saturated or undersaturated with respect to
calcium carbonate phases (e.g., calcite and aragonite);
however, as ground-water pH increases, saturation with
respect to various carbonate minerals is approached
and is usually exceeded at pH>9. This situation drives
precipitation of fairly low density minerals that over
time can impact the performance of PRBs by decreasing
reactivity (by coating iron surfaces and removing
reactive sites) and decreasing porosity and permeability
and thereby affecting hydraulic performance (U.S.
EPA, 2003a; Li et al., 2006). Geochemical modeling
results indicate that ground water upgradient of the PRB
is saturated to near-saturated with aragonite (CaCO3;
Figure 34). Ground water within the PRB is slightly
oversaturated with respect to aragonite at pH>8. These
model results and the observed decrease in calcium
concentrations between upgradient and in-wall sampling
points suggest that precipitation of aragonite is occurring
in the PRB. However, ground water entering the PRB is
sodium-sulfate type water (see Figure 10 and discussion
above). Concentrations of calcium and bicarbonate
are comparatively low (~50 mg L'1 and 200 mg I/1,
respectively) in the influent ground water. Consequently,
rapid buildup of carbonate phases is not expected.
Measurements of inorganic carbon concentrations
accumulated on the reactive medium after 15 months
range from 0.01 to 0.46 wt% (mean 0.09 wt%, n=29,
Figure 35), with the highest values observed near the
leading edge of the barrier and at shallow depths where
pH is elevated. The maximum carbonate concentration
observed after 15 months corresponds to an equivalent
reduction in the fractional porosity of <0.05 (U.S. EPA,
2003a).
c
o
O)
«
-2-
0
O O '
>°° cPoo
o o
oversaturated
undersatu rated
Upgradient aquifer
PRB
8 10
PH
12
Figure 34. Aragonite saturation indices in ground water
as a function of pH upgradient and within the
PRB.
-------
16-1
0
0.00 0.02 0.04 0.06 0.08 0.10 0.12 0.14 0.16
20-
5" 1
I '°-
,i- 5 -
Total Sulfur. wt%
100 200 300 400
Arsenic, mg/kg
500
Figure 35. Solid-phase concentrations of inorganic
carbon, sulfur, and arsenic in PRB core
materials.
Elevated concentrations of sulfate (-530 mg I/1)
in ground water entering the PRB coupled with
moderately to highly reducing conditions in the PRB
have led to the reduction of sulfate and production of
sulfide, which has precipitated in the PRB, likely as
an iron sulfide. Evidence for sulfate reduction in the
PRB includes decreases in sulfate concentrations (up
to 100%; Figure 36), detection of dissolved sulfide
within and downgradient of the PRB (up to 13 mg I/1),
and accumulation of sulfur on the reactive medium.
Reduction in sulfate concentrations is greatest at
shallower depths where residence time in the PRB is the
greatest (Figure 36 inset). Total sulfur values determined
in core samples range from 100 to 1200 mg kg"1 (mean
400 mg kg'1, n=43, Figure 35). Previous studies have
shown that sulfur accumulation in zerovalent iron PRBs
is mainly due to precipitation of mackinawite (FeS,
e.g., Roh et al., 2000; Furukawa et al., 2002; Wilkin et
al., 2003). The sulfate reduction process is microbially
mediated. Sulfate-reducing bacteria in this system
utilize either hydrogen (formed during the anaerobic
iron corrosion) or possibly organic carbon (derived
from broken-down guar gum), or both. Column and
batch experiments performed during the design phase
of the study were abiotic; sulfate was present in the
systems but behaved as a conservative tracer and was not
reduced to sulfide. Consequently this important aspect
of PRB behavior is typically not captured in laboratory
experimentation (but see Nooten et al., 2008).
1 2-
10-
08-
0.6-
0.4-
02-
00-
, ! 1
' _ :
(
EMnftM. m AUSL
B
o
fl
8
8 |
8
8 o
8
H
u O
c
8
0
§
O
8
8
8
o
o
M
0 5 10 15 20 26 30
Months of Operation
Figure 36. Sulfate removal within the PRB as a function
of time and depth. Sulfate concentration
entering the PRB, C°, is 530 mg I/1.
The role of arsenic-sulfur interactions in zerovalent
iron systems has been previously noted by Ramaswami
et al. (2001), Nikolaidis et al. (2003), and Kober et al.
(2005). Reduction of sulfate to sulfide can impact: 1) the
aqueous speciation of arsenic by driving the formation
of thioarsenic species and, 2) the removal process of
arsenic in the reactive medium. Aqueous thioarsenic
formation was noted within the PRB in several wells
with sulfide concentrations above 0.2 mg I/1. Figure 37
shows an IC-ICP-MS chromatogram for ground water
from influent well EPA08 and PRB well TR9. As noted
previously, arsenite dominates the aqueous speciation
of arsenic in influent ground water. In well TR9, the
dissolved arsenic and sulfide concentrations were
0.38 mg I/1 and 0.42 mg I/1, respectively. Arsenite
is present in trace amounts. Note that the arsenate
concentration in TR9 is similar to the influent and two
additional As-S species are present. The additional
species are thioarsenic ions that contain S and As in the
ratio of 1:1 and 2:1, respectively (Figure 37).
-------
Sxltf-
16 4x10' -
[2 3x10' -
5 2x10' -
1x10*-
0-
1
1
— EPA08
£
a
I
TR9
£
i
1 1
o
i
CO
i
pj
L^
0 2 4 S 6 10 12 14 16 18
Time, minutes
Figure 37. Chromatograph of arsenic speciation for
ground water entering the PRB (well EPA08)
and ground water from well TR9 containing
thioarsenic species.
Because arsenic and sulfur form bonded ions in the
aqueous phase, it is reasonable to expect solid-phase
associations of arsenic and sulfur in the reactive
medium. The nature of arsenic bonding in the solid-
phase is examined in a following section describing
x-ray absorption spectroscopy results. The problem is
explored here with the aid of geochemical modeling.
Removal of arsenic via the precipitation of an As(III)-
bearing phase is considered on Figure 38 showing the
pH-dependent solubility of As2O3 and As2S3. Claudetite
(As2O3) is highly soluble and this phase is not expected
to play a role in arsenic removal. Solubility of orpiment
(As2S3) is highly pH dependent and influenced by the
concentration of dissolved sulfide at pH<10. Based
on solubility reasoning, at pH>8 it is unlikely that
precipitation of As2S3 is an important arsenic removal
mechanism in zerovalent iron PRBs (Kober et al.,
2005). However, at lower pH it is possible that orpiment
precipitation could play a role in arsenic uptake. This
modeling does not rule out arsenic sorption to iron
monosulfides or thioarsenic sorption to iron corrosion
products (e.g., Gallegos et al., 2007; Wolthers et al.,
2005). Removal of arsenate via precipitation of calcium
and ferrous salts is examined on Figure 39. Note the
contrasting pH dependencies of arsenite and arsenate
precipitation. Arsenate removal is favored with
increasing pH, whereas arsenite is expected to become
increasingly more soluble at higher pH. Although
there is some uncertainty in the thermodynamic
data for symplesite (ferrous arsenate), the modeling
results suggest that symplesite, Fe3(AsO4)2-8H2O, may
precipitate in the PRB at typical conditions.
Figure 38. Solubility of As(III) phases as a function of
pH and dissolved sulfide concentration. Data
points are measured As(III) concentrations
within the PRB.
Figure 39. Solubility of As(V) phases as a function
of pH. Data points are measured As(V)
concentrations within the PRB (blue circles,
Fe=0.1 mM; red circles, Fe=10 um; green
circles, Fe=l um).
Arsenic is obviously the main contaminant of concern
for which the PRB performance is critical at this site.
Several other metals are present in the ground water
entering the PRB, including Zn (2300 |ag I/1), Cd
(50 |ag L-1), Co (20 |ag I/1), and Ni (10 |ag I/1).
Concentrations of Cd, Co, and Ni within ground water
from the PRB are below detection (generally < 1 M-g L'1),
and Zn concentrations have decreased to <50 |ag L'1.
All of these metals form insoluble precipitates with
sulfide, so that metal sulfide precipitation is one possible
removal mechanism for these contaminants, as is
adsorption onto iron corrosion products.
-------
Data for selenium are shown on Figure 40. The
histogram shows the concentration values of selenium
observed within the PRB (after 25 months) as compared
to the concentration observed in upgradient well EPA08,
which is about 12 (o,g I/1. Selenium concentrations
within the PRB span from about 5 (o,g I/1 to 18 (o,g I/1
and are log-normally distributed about the upgradient
value. These data are difficult to interpret in terms of Se
removal by the PRB. While there is a greater weighting
of observations below the upgradient concentration
within the wall, it is equally clear that some wells within
the PRB have greater concentrations than the upgradient
location.
Se values wilhin PRB
upgradient Se
concentration (12 ii
0 2 4 E 8 10 t2 14 16 18 20
Selenium. |ig/L
Figure 40. Concentration of selenium in ground water
upgradient, in, and downgradient of the PRB.
Scanning Electron Microscopy
Microscopy studies were conducted to examine
the nature of mineral deposits on the iron surfaces
that developed after 15 months, both in terms of
compositional and morphological properties. Ten
samples were selected for analysis based on location
and bulk chemical properties. Figure 4 la shows a
backscattered image of iron grains from a polished
thin-section mount. In some areas mineral precipitates
coat the iron grains; in other areas the iron grains
remain free of corrosion products. A closer view of
precipitates formed on an iron surface is shown in
Figure 41b (unpolished mount). Platy particles, rich in
iron, silicon, and oxygen, are fairly typical corrosion
products formed on granular zerovalent iron (e.g., Roh
et al., 2000; Kamolpornwijit et al., 2004; Kohn et al.,
2005). Arsenic was not detected in any of the energy
dispersive x-ray scans (n~100). The approximate
minimum detection limit for arsenic using SEM/EDX
is 0.5 wt% or 5000 mg kg"1. This concentration is
about lOx greater than the maximum bulk arsenic
concentration determined from the core materials by
acid digestion (see Figure 35). It was initially expected
that arsenic might be concentrated in the corrosion
products to detectable levels by SEM/EDX, but
exhaustive attempts to locate regions with measureable
arsenic were unsuccessful. This failure to identify
regions with concentrated arsenic concentrations may
indicate that arsenic uptake is widely distributed on the
reactive medium. The maximum arsenic concentration
determined from the cores collected after 15 months
represents about 10% of the maximum uptake capacity
determined in laboratory column tests (Lien and Wilkin,
2005). The nature of arsenic bonding to the reactive
medium is more fully explored in the next section on
x-ray absorption spectroscopy applications.
-------
0 1 2 3 4 5 6 7 8 9 It] 11 12 13 14 15 18 1718 1fl 20
Energy (keV)
Weight %
Sigma
Figure 41. SEM photomicrographs. A) Image of polished thin-section showing development of corrosion products. B)
Platy particles formed on the surface of an iron granule with EDX spectra and elemental composition.
-------
X-ray Absorption Spectroscopy
XANES analysis was conducted on reference materials,
PRB core samples, source zone materials, and well
boring samples collected in the aquifer near the location
of the pilot-PRB. These analyses were conducted to
determine the oxidation state and bonding environment
of arsenic in these different environments.
XANES spectra for the reference materials are shown
on Figure 42 and the corresponding absorption edge
and white-line positions are listed in Table 2. Reference
materials were used to analyze XANES spectra of the
unknown samples shown on Figure 43. Examination of
Figure 43 indicates no contribution of elemental arsenic
to any of the unknown samples. For this reason the
reference line for elemental arsenic has been removed
from Figure 43. Panels A and B show XANES data
for PRB core materials collected in September 2006.
Panel C shows XANES data for cores taken near
the arsenic source zone in August 2006, and Panel
D provides data for well borings collected prior to
installation of the PRB in November 2004.
ABC D
Ill
"=3"
ft
73
N
• Elemental As
•Orpi merit
As(lll) Sorted FeS
• As(lll) Sorbed Ferrihydrite
Arsertte
• Enargte
• Arsenate
• As(V) Sorbed Ferrihydrite
Scorodrte
11850
11890
Figure 42. Normalized XANES spectra for arsenic reference materials used for the XANES analysis and LCF
fitting of unknown samples. (A) is the absorption edge position for elemental As (11866.7 eV); (B) is the
absorption edge position for As(III)-S bonds (11868.7 eV); (C) is the absorption edge position for As(III)-O
bonds or As(V)-S bonds (-11870.1 eV); and, (D) is the absorption edge position for As(V)-O bonds
(11874.0 eV). Orpiment (As2S3) is used as a model As(III)-S compound; FeS is disordered mackinawite
(FeS); enargite (Cu3AsS4) is used as a model As(V)-S compound and scorodite (FeAsO4) is used as a model
forAs(V)-O.
-------
abc
Core33842F
O- Core33842
Core23845F
Core23842F
ore14448
O Core33038A
Core43034B
Core23034B
-O-Core 13034
11850 11660 11870 11880 11860 11850 11
11870 11880 11890
APBH4210115
-O-APBH41565
TW183840
- •- TW162530
-O-TW152025
-O-9.1 m
O 10.7m
E(eV)
Figure 43. Normalized XANES spectra for unknown samples from the ASARCO Smelter site. PRB core sample
spectra are shown in Panels A (shallow depths) and B (deep depths). Panel C shows spectra for samples
collected from the source zone and Panel D shows spectra of samples from a well boring taken prior to
installation of the PRB. (a) is the white-line position for As(III)-S, (b) is the white-line for the As(III)-O,
and (c) is the white-line position for the As(V)-O.
XANES spectra from the PRB samples show that
there are multiple solid-phase arsenic species present.
Arsenic in all the PRB samples includes at least two
species: As(III)-O and As(V)-O, but in many samples
As(III)-S appears to be present as well. It should be
noted that As(V)-S cannot be ruled out in any of the
samples because the As(III)-O and As(V)-S white line
positions overlap (Beak et al., 2008), although this
solid-phase coordination is considered to be unlikely
in this environment. The source zone samples appear
to contain only As(V)-O. The presence of As(III)
oxyanions in the source zone ground water suggests the
possibility of microbial reduction of arsenate to arsenite
in the subsurface.
Multiple arsenic species are also present in well
borings collected from the aquifer near the location
of the pilot-PRB. In shallow samples the speciation
appears to be an unusual mixture of As(V)-O and
As(III)-S (Figure 43a). To our knowledge this unique
arsenic speciation has not been observed in other
subsurface environments. However, the observed
As(III)-S speciation can be explained by the presence
of a petroleum hydrocarbon plume that induces
microbially mediated redox reactions. Degradation of
the hydrocarbon leads to local iron- and surf ate-reducing
conditions that result in the formation of Fe-S surfaces
and As(III)-S speciation at shallow depths. Deeper in
the aquifer, arsenic speciation is a more typical mixture
-------
of As(III)-O and As(V)-O. Figure 44 shows the depth-
dependent arsenic concentration and speciation profile
through the unsaturated zone and saturated aquifer near
the location of the PRB. In the unsaturated zone, total
arsenic concentrations are low (<25 mg kg"1) and the
solid-phase speciation is completely as As(V). Maximal
arsenic concentrations are observed in the reducing zone
near the water table, indicating that arsenic attenuation
can occur in iron- and sulfate-reducing environments.
A profile of x-ray diffraction scans was also collected
for the core samples (Figure 45). The mineralogical
analysis indicates the presence of a mixture of quartz,
feldspars, and clay minerals. Any reduced, sulfide-
bearing minerals are not concentrated enough to be
detectable by bulk XRD methods.
1190
1 iss-
ues
U> 1184
E 1182-
» 1180-
1
1178-
1176-
1174
Land surface
a)
1190
1188-
1186-
1184-
1182-
1180-
1178-
1176-
1174
0 100 200 300 400 500 600 700 800 900 0.0 0.5 1.0
Arsenic, mg/kg fraction As(V)
Figure 44. a) Depth-dependent concentrations of total
arsenic in aquifer solids retrieved from the
well boring near the location of the PRB,
and b) speciation of arsenic expressed as the
fraction of total arsenic as As(V) in the solid
samples.
-njuuu_
Depth below ground surface
3.0 m
7.6 m
9.1 m
10.7m
13.7m
14.7m
Quartz
10 20 30 40 50 60 70 80 90
°29
Figure 45. Powder x-ray diffraction scans of aquifer materials collected from a well boring taken adjacent to the pilot-
PRB.
-------
XANES analysis of unknown samples by comparison
of white line and absorption edge positions to reference
materials demonstrates that XANES alone is not
sensitive enough to distinguish between arsenic species
(see Figure 42 and 43). Because of this limitation
linear combination fitting (LCF) was employed to
further analyze the XANES data. LCF consisted of
using reference spectra of several possible solid-phase
arsenic species to allow the "quantification" of species in
unknown, multiple-component mixtures. This method
uses a least-squares fitting algorithm to refine the sum
of reference spectra to an experimental spectrum. The
reference spectra shown on Figure 42 were used in the
LCF fitting of the unknown samples. Examination of
Figure 42 and Table 1 shows that the reference materials
fall into two general categories: 1) arsenic sorbed to
mineral phases (sorbed), and 2) pure arsenic minerals
that contain As(III) or As(V) coordinated to S or O
(coordinated). LCF analysis was carried out using both
types of reference materials and results are shown in
Table 6. The data in Table 6 illustrate that in most cases
the values determined for solid-phase species were
similar, whether coordinated or sorbed models were
adopted. In fact, ANOVA analysis of the spectral fit
data indicates no statistical difference between sorbed
and coordinated models. Fitting statistics (R-Factors)
of the sorbed model were significantly better than the
coordinated model as shown visually on Figure 46.
The improved fit suggests that the arsenic species present
in the samples are primarily sorbed to iron oxide or
sulfide surfaces. We cannot rule out the presence of pure
arsenic phases, although LCF analysis does not support
these species being present in significant amounts.
Table 6. Results of LCF fitting of unknown samples collected from the Asarco Smelter.
Coordinated
Core
PRB1
PRB1
PRB2
PRB2
PRB2
PRB3
PRB3
PRB3
PRB4
APBH 4-1
APBH 4-2
TW1-5
TW1-6
TW1-8
Boring
Boring
Boring
Boring
Depth (ft)
30-34F
44-48
30-34B
38-42F
38-45F
30-38A
38-42
38-42F
30.34B
1.5-6.5
10-11.5
20-25
25-30
38-40
30
35
45
50
As(V)-O
28.8
18.5
35.9
14.4
22.4
37.3
26.1
20.1
39.9
93.5
72.0
50.0
65.8
73.4
28.7
30.2
50.3
51.1
fitting model
As(III)-O
41.7
55.1
39.3
60.5
60.6
37.3
48.1
29.1
45.8
6.5
28.0
27.0
34.2
26.6
0.9
8.5
49.7
48.9
As(III)-S
29.4
26.4
24.8
25.1
17.1
25.4
25.8
50.8
14.3
0.0
0.0
23.0
0.0
0.0
70.4
61.2
0.0
0.0
R-Factor
coordinated
0.004
0.015
0.006
0.011
0.011
0.019
0.017
0.011
0.005
0.015
0.015
0.011
0.015
0.011
0.009
0.011
0.014
0.016
Sorbed fitting model
As(V)-Fh
31.7
19.3
40.6
19.1
26.6
41.8
29.5
16.6
48.4
99.4
79.8
57.0
75.4
81.8
18.0
22.7
63.9
64.1
As(III)-Fh
30.4
59.3
31.9
54.1
62.2
41.8
35.9
18.7
26.2
0.6
20.2
43.0
24.6
18.2
0.0
0.0
28.7
35.4
* /TTTX T- o R-Factor
As(III)-FeS , ,
sorbed
37.9
21.4
27.5
26.9
11.2
16.4
34.6
64.7
25.3
0.0
0.0
0.0
0.0
0.0
82.0
77.3
7.4
0.5
0.0010
0.0013
0.0015
0.0008
0.0008
0.0060
0.0037
0.0013
0.0017
0.0060
0.0050
0.0015
0.0050
0.0030
0.0030
0.0020
0.0050
0.0060
The results of both fitting types, minerals containing As(III) or As(V) coordinated with S or O and As(III) or As(V) sorbed to
mineral surfaces, are shown in this table. As(V)-O is As(V) coordinated with O as in arsenate; As(V)-Fh is As(V) sorbed to
ferrihydrite; As(III)-O is As(III) coordinated with oxygen as in arsenite; As(III) is As(III) sorbed to ferrihydrite; As(III)-S is
As(III) coordinated with S as in the mineral orpiment; and As(III)-FeS is As(III) sorbed to FeS. The R-Factor is a statistical
measure of the error of the fit. As(V)-S was not found to contribute to the fits in any of the samples and was therefore dropped
from the table. Also, there was no standard reference material for As(V) sorbed to FeS used in the fitting.
-------
ty.
a
11850 11860 11870 11880 1189011850 11860 11870 11860 11890
E (eV) E (eV)
Figure 46. Comparison of LCF fitting results using a)
coordinated or b) sorbed reference materials
forPRB core sample Core 1 30-34F. The
solid black line is the measured spectra and
the open red circles are the fits.
Keeping the previous discussion in mind and using
the LCF and XANES data we can make some
generalizations about the speciation of arsenic in the
solid samples collected from the source zone, aquifer
adjacent to the PRB, and the PRB. A ternary diagram is
shown in Figure 47 that shows the relative proportions
of As(III)-O, As(V)-O, and As(III)-S bonding
environments. In all cases the PRB core samples likely
contained three species, As(V) and As(III) sorbed to Fe
(oxy)hydroxides and As(III) sorbed to Fe sulfide phases.
Samples collected in the source zone contained two As
species As(V) and As(III) sorbed to Fe (oxy)hydroxides
and in the case of sample APBH 4-1 the speciation was
primarily As(V) sorbed to Fe (oxy)hydroxides, although
there was a small amount of As(III) sorbed to (oxy)
hydroxides. The well borings collected prior to PRB
installation show that the speciation in the shallow
depths was primarily As(III) sorbed to FeS, with a small
fraction of the speciation being As(V) sorbed to Fe (oxy)
hydroxides. Again this is likely because of hydrocarbon
contamination driving microbial processes that result
in iron- and sulfate-reducing conditions. Finally, in
the deeper depths the speciation was As(V) and As(III)
sorbed to Fe (oxy)hydroxides.
PRB 0.00^1.00
Source Zone
Borings
0.25,
0.50.
1.00
0.75
0.00
0.00
0.25
0.50
As(V)-O
0.75
1.00
Figure 47. Ternary diagram showing the solid-phase arsenic speciation based on XANES LCF in samples collected
from the PRB, source zone, and aquifer adjacent to the pilot-PRB.
-------
7.0
Future Study Improvements
At this point it may be worthwhile to point out areas
where this study could have been improved to help
future efforts of this sort and to indicate planned
continued studies. One obvious shortcoming of the
study is the fact that the lower confining unit was not
tagged during construction of the pilot PRB. As noted in
the report, however, the negative consequences exceeded
the possible benefit. More information is needed at this
site regarding the continuity and thickness of the lower
ash confining unit. The initial high DOC concentrations
were a bit surprising and were not helpful in the early
ground-water sampling campaigns. While it is possible
that microbial populations were stimulated by this dose
of organic carbon, it may have been beneficial to spend
more time and effort in attempts to actively degrade the
residual bioslurry. Future plans involve more detailed
tracer studies that combine heat and chemical tracers.
Spectroscopic studies that move beyond the near-edge
region of arsenic are also planned on a new set of cores
to be collected after about 3.5 y of operation. These
cores will be subjected to XRD analysis, as well as As
EXAFS and Fe EXAFS analysis, to better fingerprint the
important arsenic uptake and iron corrosion processes.
-------
8.0
Summary and Relevance to Other Sites
In June 2005, a 9.1 m long, 14 m deep, and 1.8 to 2.4 m
wide (in the direction of ground-water flow) pilot-scale
permeable reactive barrier (PRB) was installed at a
former metal smelting facility, located near Helena,
Montana. The reactive barrier was installed over a 3-day
period using bio-polymer slurry methods and modified
excavating equipment for deep trenching. The reactive
medium was composed entirely of granular iron. A
monitoring network of approximately 40 ground-water
sampling wells was installed in July 2005. Ground
water samples were collected and analyzed at 1 month,
4 months, 12 months, 15 and 25 months of operation.
After over 2 years of monitoring, results indicate
arsenic concentrations >25 mg I/1 in wells located
hydraulically upgradient of the PRB. Within the PRB,
arsenic concentrations are reduced to 2 to <0.01 mg I/1.
Detailed studies in the aquifer downgradient of the PRB
show an upper zone of the saturated aquifer (8.8 m to
12.8 m below land surface) where arsenic concentrations
are reduced to <0.5 mg I/1 after 25 months of operation.
Arsenic concentrations in the lower zone of the
downgradient region, from 12.8 m to 14.6 m below land
surface, increase with increasing depth to a maximum
value of about 27 mg I/1, or roughly the same arsenic
concentration observed on the upgradient side of the
PRB. Ineffective treatment of arsenic over the lower
depth interval is likely due to the fact that the maximum
PRB depth is 13.7-14.0 m below land surface, about
1 m short of the depth of the basal weathered ash. The
pilot-PRB was not anchored into the weathered ash
because of uncertainties about the ash thickness in the
area of the site where the PRB was installed and to
insure that the lower confining unit was not breached
during the excavation. In short, where hydraulic
connection between the upgradient aquifer and the PRB
is established, the pilot PRB is performing as expected.
Monitoring of the water table upgradient of the PRB
indicates that mounding of groundwater on the leading
edge of the barrier is negligible. In addition, slug testing
in the PRB indicates hydraulic conductivities average
approximately 60 m d1, a value that is comparable to the
native aquifer materials in this region of the site. These
data, along with other geochemical indicators (such as
reduced arsenic concentrations in the PRB), indicate
that the system is meeting hydraulic performance
expectations.
General conclusions and observations can be made
regarding the application of PRB technology with
granular iron for treatment of arsenic in ground water.
• Zerovalent iron can be effectively used to treat ground
water contaminated with arsenic given appropriate
geochemical and hydrological conditions. Results
of pilot testing are promising in that significant
concentration reductions for arsenic have been
achieved for over 2 years in the downgradient aquifer
directly impacted by the PRB and the system is
meeting hydraulic performance expectations. Solid-
phase tests indicate that <10% of the arsenic uptake
capacity of the reactive medium was used after about
1 year of operation.
• Arsenic removal processes appear to be complex as
multiple bonding and coordination environments
around arsenic are revealed using advanced
spectroscopic techniques. Sequestration mechanisms
taking place in the field are more varied compared
to those previously indicated in highly controlled
laboratory tests. PRB core samples likely contain
arsenic in three solid-phase species: As(V) and As(III)
sorbed to Fe (oxy)hydroxides and As(III) sorbed to Fe
sulfide phases. This study demonstrated a relationship
between observed distributions of aqueous species of
arsenic with solid-phase species as determined using
spectroscopic analyses.
• Granular iron has a finite capacity to remove arsenic
from solution. Therefore, successful applications
of PRB technology for arsenic removal require
detailed subsurface characterization data that
capture geochemical and hydrogeologic variability.
Evaluation of depth-dependent arsenic flux is critical.
Zones with maximal contaminant flux will necessarily
dictate system design requirements. This study
employed discrete interval sampling for geochemical
profiles, interval-specific probes of hydraulic
conductivity, and continuous water level logging to
evaluate long-term arsenic transport in the subsurface.
Flux evaluations indicate that arsenic removal by
zerovalent is highly effective at loadings below
5 g As/m2d.
• Sodium-sulfate-type ground water is better suited
for contaminant treatment by zerovalent iron over
calcium-bicarbonate-type ground water. Carbonate
-------
precipitation is an undesirable consequence of • Source control measures will reduce arsenic loading to
zerovalent iron applications and is expected to reduce the PRB and implementation of such measures should
long-term reactive and hydraulic performance. result in increased PRB lifetimes. The increase in
Microbial reduction of sulfate to sulfide and effective lifetime should be directly proportional to the
precipitation of low-density iron sulfides within the level of concentration reduction achieved by isolating
reactive medium creates additional mineral surfaces and/or treating the source of contamination.
for arsenic removal.
-------
9.0
References
Bain, J., D. Blowes, D. Smyth, C. Ptacek, J. Wilkens,
and R. Ludwig. "Permeable reactive barriers for the
in-situ treatment of arsenic contaminated groundwater."
In Proceedings of the 5th International Conference
on Remediation of Chlorinated and Recalcitrant
Compounds, Battelle Press, Columbus, OH, 2006.
Bang, S., M. D. Johnson, G. P. Korfiatis, and X. Meng.
Chemical reactions between arsenic and zero-valent iron
in water. Water Research 39: 763-770 (2005a).
Bang, S., G. P. Korfiatis, and X. Meng. Removal of
arsenic from water by zero-valent iron. Journal of
Hazardous Materials 121: 61-67 (2005b).
Beak, D., R. T. Wilkin, R. G. Ford, and S. Kelly. An
examination of arsenic speciation in sulfidic solutions
using x-ray absorption spectroscopy. Environmental
Science and Technology 42: 1643-1650 (2008).
Blowes, D. W., C. J. Ptacek, S. G. Benner, C. W. T.
McRae, T. A. Bennett, and R. W. Puls. Treament
of inorganic contaminants using permeable reactive
barriers. Journal of Contaminant Hydrology 45: 123-137
(2000).
Biterna, M., A. Arditsoglou, E. Tsikouras, and D. Voutsa.
Arsenate removal by zero valent iron: Batch and column
tests. Journal of Hazardous Materials 149: 548-552
(2007).
Bouwer, H. and R. C. Rice. A slug test method for
determining hydraulic conductivity of unconfmed
aquifers with completely or partially penetrating wells.
Water Resources Research 12: 423-428 (1976).
Briar, D. W. and J. P. Madison. Hydrogeology of the
Helena valley-fill aquifer system, west-central Montana.
Water-Resources Investigations Report 92-4023, U.S.
Geological Survey, Reston, VA, 1992.
Butler, J. J., Jr. The Design, Performance, and Analysis
of Slug Tests, Lewis Publishers, Boca Raton, 1997.
Butler, J. J., Jr. and E. J. Garnett. Simple procedures for
analyses of slug tests informations of high hydraulic
conductivity using spreadsheet and scientific graphics
software. Kansas Geological Survey, Open File Report
2000-40, 2000.
Cherry, J. A., A. U. Shaikh, D. E. Tallman, and R. V
Nicholson. Arsenic species as an indicator of redox
conditions in groundwater. Journal of Hydrology 43:
373-392 (1979).
Cullen, W. R. and K. J. Reimer. Arsenic speciation in the
environment. Chemical Reviews 89: 713-764 (1989).
Del Razo, L. M., M. A. Arellano, and M. E. Cebrian,.
The oxidation states of arsenic in well water from
a chronic arsenicism area of northern Mexico.
Environmental Pollution 64: 143-153 (1990).
Devlin, J. F. A spreadsheet method of estimating best-fit
hydraulic gradients using head data from multiple wells.
Ground Water 41: 316-320 (2003).
Eary, L. E. The solubility of amorphous As2S3 from 25
to 90 °C. Geochimica et Cosmochimica Acta 56: 2267-
2280 (1992).
Farrell, J., J. Wang, P. O'Day, and M. Conklin.
Electrochemical and spectroscopic study of arsenate
removal from water using zero-valent iron media.
Environmental Science and Technology 35: 2026-2032
(2001).
Ferguson, J. F. and J. Gavis. A review of the arsenic
cycle in natural waters. Water Research 6: 1259-1274
(1972).
Furukawa, Y, J. Kim, J. Watkins, and R. T. Wilkin.
Formation of ferrihydrite and associated corrosion
products in Permeable Reactive Barriers of zero-valent
iron. Environmental Science and Technology 36: 5469-
5475 (2002).
Gallegos, T. J., S. P. Hyun, and K. F. Hayes.
Spectroscopic investigation of the uptake of arsenite
from solution by synthetic mackinawite. Environmental
Science and Technology 41: 7781-7786 (2007).
Holm, T. R. and C. D. Curtiss. A comparison of
oxidation-reduction potentials calculated from the
As(V)/As(III) and Fe(III)/Fe(II) couples with measured
platinum-electrode potentials in groundwater. Journal of
Contaminant Hydrology 5: 67-81 (1989).
ITRC. Permeable Reactive Barriers: Lessons learned/
new directions. Interstate Technology and Regulatory
Council, Report PRB-4, Washington, DC, 2005.
-------
Johnston, R. B. and P. C. Singer. Redox reactions in the
Fe-As-O2 system. Chemosphere 69: 517-525 (2007a).
Johnston, R. B. and Singer, P. C. Solubility of
symplesite (ferrous arsenate): implications for reduced
groundwaters and other geochemical environments.
5*0/7 Science Society of America Journal 71: 101-107
(2007b).
Kamolpornwijit, W., L. Liang, G. R. Moline, T. Hart,
and O. R. West. Identification and quantification of
mineral precipitation in FeO filings from a column study.
Environmental Science and Technology 38: 5757-5765
(2004).
Keys, W. S. and R. F. Brown. The use of temperature
logs to trace the movement of injected water. Ground
Water 16: 32-48 (1978).
Kober, R., E. Welter, M. Ebert, and A. Dahmke.
Removal of arsenic from groundwater by zerovalent
iron and the role of sulfide. Environmental Science and
Technology 39: 8038-8044 (2005).
Kohn, T., K. J. T. Livi, A. L. Roberts, and P. J.
Vikesland. Longevity of granular iron in groundwater
treatment processes: Corrosion product development.
Environmental Science and Technology 39: 2867-2879
(2005).
Lackovic, J. A., N. P. Nikolaidis, and G. M. Dobbs.
Inorganic arsenic removal by zero-valent iron.
Environmental Engineering Science 17: 29-39 (2000).
Leupin, O. X. and S. J. Hug. Oxidation and removal
of arsenic(III) from aerated groundwater by filtration
through sand and zero-valent iron. Water Research
39: 1729-1740 (2005).
Li, L., C. H. Benson, and E. M. Lawson. Modeling
porosity reductions caused by mineral fouling in
continuous-wall permeable reactive barriers. Journal of
Contaminant Hydrogeology%?>: 89-121 (2006).
Liang, L., G. R. Moline, W. Kamolpornwijit, and O.
R. West. Influence of hydrochemical processes on
zero-valent iron reactive barrier performance: A field
investigation. Journal of Contaminant Hydrogeology
80: 71-91 (2005).
Lien, H.-L. and R. T. Wilkin. High-level arsenite
removal from groundwater by zero-valent iron.
Chemosphere 59: 377-386 (2005).
Lorenz, H. W. and F. A. Swenson. Geology and ground-
water resources of the Helena Valley, Montana. Circular
83, U.S. Geological Survey, Reston, VA, 1951.
Magalhaes, M. C. F. Arsenic. An environmental
problem limited by solubility. Pure and Applied
Chemistry'74: 1843-1850 (2002).
Mandal, B. Chronic arsenic toxicity in West Bengal.
Current Science 72: 917-924 (1997).
Manning, B. A., M. L. Hunt, C. Amrhein, and J. A.
Yarmoff. Arsenic(III) and arsenic(V) reactions with
zerovalent iron corrosion products. Environmental
Science and Technology 36: 5455-5461 (2002).
Me Arthur, J. M., P. Ravenscroft, S. Safiulla, and M. F.
Thirlwall. Arsenic in groundwater: Testing pollution
mechanisms for sedimentary aquifers in Bangladesh.
Water Resources Research 37: 109-117 (2001).
Melitas, N., J. Wang, M. Conklin, P. O'Day, and J.
Farrell. Understanding soluble arsenate removal kinetics
by zerovalent iron media. Environmental Science and
Technology 36: 2074-2081(2002).
Molz, F. J. and S. C. Young. Development and
application of borehole flowmeters for environmental
assessment. The Log Analyst 34: 13-23 (1993).
Molz, F. J., G. K. Boman, S. C. Young, and W. R.
Waldrop. Borehole flowmeters: field application and
data analysis. Journal of Hydrology 163: 347-371
(1994).
Morrison, S. J., D. R. Metzler, and B. P. Dwyer.
Removal of As, Mn, Mo, U, V, and Zn from groundwater
by zero-valent iron in a passive treatment cell: reaction
progress modeling. Journal of Contaminant Hydrology
56:99-116(2002).
Morrison, S. J., P. S. Mushovic, and P. L. Niesen.
Early breakthrough of molybdenum and uranium in a
permeable reactive barrier. Environmental Science and
Technology 40: 2018-2024 (2006).
National Research Council. Arsenic in Drinking Water.
National Academy Press, Washington, DC, 1999.
Newville, M. IFEFFIT: Interactive EXAFS analysis and
FEFF fitting. Journal of Synchrotron Radiation 322-324
(2001).
Nickson, R., J. McArthur, W. Burgess, K. M. Ahmed,
P. Ravenscroft, and M. Rahman. Arsenic poisoning of
Bangladesh groundwater. Nature 395: 338 (1998).
Nikolaidis, N. P., G. M. Dobbs, and J. A. Lackovic.
Arsenic removal by zero-valent iron: field, laboratory
and modeling studies. Water Research 37': 1417-1425
(2003).
-------
Nooten, T. V, D. Springael, and L. Bastiaens. Positive
impact of microorganisms on the performance of
laboratory-scale permeable reactive iron barriers.
Environmental Science and Technology 42: 1680-1686
(2008).
Nordstrom, D. K. Worldwide occurrences of arsenic in
ground water. Science 296: 2143-2145 (2002).
Nordstrom, D. K. and D. G. Archer. "Arsenic
thermodynamic data and environmental geochemistry."
In Arsenic in Ground Water: Geochemistry and
Occurrence, A. H. Welch and K. G. Stollenwerk (eds.),
Kluwer Academic Publishers, The Netherlands, 2003.
O'Hannesin, S. and R. W. Gillham. Long-term
performance of an in situ "iron wall" for remediation of
VOCs. Ground Water 36: 164-170 (1998).
Polizzoto, M. L., C. F. Harvey, G. Li, B. Badruzzman,
A. Ali, M. Newville, S. Sutton, and S. Fendorf. Solid-
phases and desorption processes of arsenic within
Bangladesh sediments. Chemical Geology 228: 97-111
(2006).
Rader, K., P. M. Dombrowski, K. J. Farley, J. D.
Mahony, and D. M. Ditoro. Effect of thioarsenite
formation on arsenic(III) toxicity. Environmental
Toxicology and Chemistry 23: 1649-1654 (2004).
Rahman, M., U. Chowdhury, S. Mukherjee, B. Mandal,
K. Paul, D. Lodh, B. Biswas, C. Chanda, G. Basu, K.
Sana, S. Roy, R. Das, S. Palit, Q. Quamruzzaman, and
D. Chakraborti. Chronic arsenic toxicity in Bangladesh
and West Bengal, India-A review and commentary.
Toxicology and Clinical Toxicology 39: 683 (2001).
Ramaswami, A., S. Tawachsupa, and M. Isleyen. Batch-
mixed iron treatment of high arsenic waters. Water
Research 35: 4474-4479 (2001).
Ravel, B. and M. Newville. Athena, Artemis,
Hephaestus: data analysis for x-ray absorption
spectroscopy using IFEFFIT. Journal of Synchrotron
Radiation 12: 537-541 (2005).
Roh, Y, S. Y. Lee, and M. P. Elless. Characterization of
corrosion products in the permeable reactive barriers.
Environmental Geology 40: 184-194 (2000).
Ronen, D., M. Magaritz, and I. Levy. An in-situ
multilevel sampler for preventive monitoring and study
of hydrochemical profiles in aquifers. Ground Water
Monitoring Review 7: 69-74 (1986).
Rude, T. R. and S. Wohnlich. "The couple of As(V) -
As(III) as a redox indicator." In Redox: fundamentals,
processes, and applications, J. Schilling et al. (eds.).
Springer-Verlag, Berlin, 2000.
Shaw Environmental, Inc. (2005a). Technical Work Plan,
Pilot-Scale Installation of a Permeable Reactive Barrier
for the Treatment of Arsenic in Groundwater, May 2005.
Shaw Environmental, Inc. (2005b). Post Wall Installation
Report: Pilot-Scale Permeable Reactive Barrier for the
Treatment of Arsenic in Groundwater, July 2005.
Smedley, P. L. and D. G. Kinniburgh. A review of the
source, behaviour, and distribution of arsenic in natural
waters. Applied Geochemistry 17: 517-568 (2002).
Springer, R. K. and L. W. Gelhar. Characterization of
large-scale aquifer heterogeneity in glacial outwash by
analysis of slug tests with oscillatory response, Cape
Cod, Massachusetts, U.S. Geological Survey, Water
Resources Investigations Report 91-4034, 1991.
Stickney, M. C. Quaternary geology and faulting in
the Helena Valley, Montana. Geologic Map Series 46,
Montana Bureau of Mines and Geology, Butte, MT,
1987.
Su, C. and R. W. Puls. Arsenate and arsenite removal
by zerovalent iron: Kinetics, redox transformation,
and implications for in situ groundwater remediation.
Environmental Science and Technology 35: 1487-1492
(200 la).
Su, C. and R. W. Puls. Arsenate and arsenite removal by
zerovalent iron: Effects of phosphate, silicate, carbonate,
borate, sulfate, chromate, molybdate, and nitrate, relative
to chloride. Environmental Science and Technology 35:
4562-4568 (200Ib).
Su, C. and R. W. Puls. In situ remediation of arsenic in
simulated groundwater using zerovalent iron: Laboratory
column tests on combined effects of phosphate and
silicate . Environmental Science and Technology
37: 2582-2587 (2003).
Sun, H., L. Wang, R. Zhang, J. Sui, and G. Xu.
Treatment of groundwater polluted by arsenic
compounds by zero valent iron. Journal of Hazardous
Materials m29: 297-303 (2006).
USEPA. Technology alternatives for the remediation
of soils contaminated with As, Cd, Cr, Hg, andPb,
EPA/540/S-97/500. U. S. Environmental Protection
Agency, 1997.
USEPA. The QTRACER2 program for tracer-
breakthrough curve analysis for tracer tests in karstic
and other hydrologic systems, EPA/600/R-02/001.
-------
Washington, D. C.: U. S. Environmental Protection
Agency, 2002.
USEPA. Capstone report on the application,
monitoring, and performance of permeable reactive
barriers for ground-water remediation; Volume 1,
Performance evaluations at two sites, EPA/600/R-
03/045a. Cincinnati, OH: USEPA National Risk
Management Research Laboratory, 2003a.
USEPA. Capstone report on the application,
monitoring, and performance of permeable reactive
barriers for ground-water remediation; Volume 2,
Soiling sampling and ground water sampling, EPA/600/
R-03/045b. Cincinnati, OH: USEPA National Risk
Management Research Laboratory, 2003b.
USEPA. Monitored Natural Attenuation of Inorganic
Contaminants in Ground Water: Volume 2, Assessment
for Non-radionuclides Including Arsenic, Cadmium,
Chromium, Copper, Lead, Nickel, Nitrate, Perchlorate,
and Selenium, EPA Report, EPA/600/R07/140, 2007.
Vlassopoulos, D., N. Rivera, P. A. O'Day, M. T. Rafferty,
and C. B. Andrews. "Arsenic removal by zero-valent
iron: a field study of rates, mechanisms, and long
term performance." In Advances in Arsenic Research:
Integration of Experimental and Observational Studies
and Implications for Migration, PA. O'Day et al. (eds.).
American Chemical Society Symposium Series 915,
Ch. 25, 2005.
Webster, J. G. The solubility of As2S3 and speciation
of As in dilute and sulphide-bearing fluids at 25 and 90
°C. Geochimica et Cosmochimica Acta 54: 1009-1017
(1990).
Welch, A. H., M. S. Lico, and J. L. Hughes. Arsenic
in ground water of the western United States. Ground
Water 26: 333-347 (1988).
Welch, A. H., D. B. Westjohn, D. R. Helsel, and R. B.
Wanty. Arsenic in ground water of the United States:
occurrence and geochemistry. Ground Water 38: 589-604
(2000).
Wilkin, R. T., D. Wallschlager, and R. G. Ford.
Speciation of arsenic in sulfidic waters. Geochemical
Transactions 4: 1-7 (2003).
Wilkin, R. T., R. W. Puls, and G. W Sewell. Long-term
performance of Permeable Reactive Barriers using zero-
valent iron: geochemical and microbiological effects.
Ground Water 41: 493-503 (2003).
Wilkin, R. T., C. Su, R. G. Ford, and C. J. Paul.
Chromium removal processes during groundwater
remediation by a zerovalent iron permeable reactive
barrier. Environmental Science and Technology
39: 4599-4605 (2005).
Wilkin, R. T. and K. J. Bischoff. Coulometric
determination of total sulfur and reduced inorganic sulfur
fractions in environmental samples. Talanta 70: 766-773
(2006).
Wolthers, M., L. Charlet, C. H. van der Weijden, P. R.
van der Linde, and D. Rickard. Arsenic mobility in the
ambient sulfidic environment: sorption of arsenic(V) and
arsenic(III) onto disordered mackinawite. Geochimica et
Cosmochimica Acta 69: 3483-3492 (2005).
Young, S. C., H. E. Julian, H. S. Pearson, F. J. Molz,
and G. K. Boman. Application of the electromagnetic
boreholeflowmeter, EPA/600/R-98/058. Ada, OK: U.S.
Environmental Protection Agency, 1998.
Yuan, C. and T.-S. Chiang, The mechanisms of arsenic
removal from soil by electrokinetic process coupled with
iron permeable reaction barrier. Chemosphere 67: 1533-
1542 (2007).
-------
Appendix A
Table Al. Wells used in the design and assessment of the PRB.
Existing Wells:
WELL
DH-8
DH-17
DH-21
DH-24
DH-27
DH-33
DH-34
DH-36
DH-49
DH-50
DH-51
DH-56
DH-63
DH-64
DH-66
STW-1
STW-4
STW-7
STW-9
PRB1
PRB2
PRB3
Ground Surface
Elevation
(mAMSL)
1192.61
1189.34
1190.53
1187.91
1191.53
1191.43
1190.75
1189.44
1188.15
1188.35
1188.36
1204.58
1188.64
1188.20
1191.26
1188.65
1188.14
1188.67
1188.23
1190.24
1188.80
1190.96
Top of Screen
(m BLS)
11.9
9.4
5.8
8.2
5.8
6.1
6.1
6.4
7.3
7.3
7.3
21.3
7.3
13.7
11.6
10.1
9.8
7.6
10.7
10.7
11.3
11.0
Bottom of Screen
(m BLS)
14.9
12.5
8.8
10.7
8.8
9.1
9.1
9.4
10.4
10.4
10.4
25.9
11.9
16.8
14.6
11.6
11.3
12.2
12.2
15.2
15.8
15.5
Top of Screen
(mAMSL)
1180.7
1179.9
1184.7
1179.7
1185.7
1185.3
1184.7
1183.0
1180.8
1181.0
1181.0
1183.2
1181.3
1174.5
1179.7
1178.6
1178.4
1181.0
1177.6
1179.6
1177.5
1180.0
Bottom of Screen
(mAMSL)
1177.7
1176.8
1181.7
1177.2
1182.7
1182.3
1181.6
1180.0
1177.8
1178.0
1178.0
1178.7
1176.8
1171.4
1176.6
1177.1
1176.9
1176.5
1176.0
1175.0
1172.9
1175.4
Construction
4" PVC
4" PVC
4" PVC
4" PVC
4" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
Note: AMSL = Above Mean Sea Level.
BLS = Below Land Surface.
-------
Wells Installed by USEPA:
WELL
EPA01
EPA02
EPA03
EPA04
EPA05
EPA06
EPA07
EPA08
EPA09
EPA10
PBTW-1
PBTW-2
SOI
S02
803
S04
805
S06
S07
SOS
T1A
TIB
TIC
T2A
T2B
T2C
T3A
T3B
T3C
TR1
TR2
TR3
TR4-1
TR4-2
TR4-3
TR5-1
TR5-2
TR5-3
TR6-1
TR6-2
TR6-3
TR7
TR8
TR9
TRIO
TR11-1
TR11-2
TR11-3
TR12-1
TR12-2
TR12-3
Ground Surface
Elevation
(m AMSL)
1189.72
1189.59
1189.35
1189.41
1188.89
1189.45
1188.99
1189.63
1189.60
1189.57
1189.29
1189.01
1189.57
1189.57
1189.58
1189.57
1189.59
1189.57
1189.61
1189.61
1189.70
1189.72
1189.72
1189.58
1189.57
1189.59
1189.56
1189.55
1189.56
1189.61
1189.56
1189.58
1189.57
1189.57
1189.57
1189.57
1189.57
1189.57
1189.57
1189.57
1189.57
1189.59
1189.60
1189.60
1189.61
1189.60
1189.60
1189.60
1189.60
1189.60
1189.60
Top of
Screen
(m BLS)
10.1
10.1
10.1
10.1
10.1
10.1
10.1
10.1
10.1
9.1
9.4
9.8
13.0
9.9
13.0
9.9
11.4
12.2
12.2
11.4
9.6
9.4
9.4
9.2
8.8
8.8
8.6
9.3
8.7
10.4
9.9
11.4
10.0
11.5
13.0
10.0
11.5
13.1
10.0
11.5
13.0
11.4
11.4
9.9
11.4
10.0
11.5
13.0
10.0
11.6
13.1
Bottom
of Screen
(m BLS)
14.6
14.6
14.6
14.6
14.6
14.6
14.6
14.6
14.6
14.3
13.7
16.5
13.7
10.7
13.7
10.7
12.2
13.0
13.0
12.2
11.9
11.6
11.7
13.8
13.4
13.4
13.2
13.8
13.2
13.4
10.7
12.2
10.6
12.1
13.6
10.6
12.1
13.7
10.6
12.1
13.6
12.2
12.2
10.7
12.2
10.6
12.1
13.6
10.6
12.2
13.7
Top of
Screen
(mAMSL)
1179.7
1179.5
1179.3
1179.4
1178.8
1179.4
1178.9
1179.6
1179.5
1180.4
1179.8
1179.3
1176.6
1179.7
1176.6
1179.7
1178.2
1177.4
1177.4
1178.2
1180.1
1180.4
1180.3
1180.4
1180.8
1180.8
1181.0
1180.3
1180.9
1179.2
1179.7
1178.1
1179.6
1178.1
1176.6
1179.6
1178.0
1176.5
1179.6
1178.1
1176.6
1178.2
1178.2
1179.7
1178.2
1179.6
1178.1
1176.6
1179.6
1178.1
1176.5
Bottom
of Screen
(mAMSL)
1175.1
1175.0
1174.7
1174.8
1174.3
1174.8
1174.4
1175.0
1175.0
1175.2
1175.6
1172.6
1175.9
1178.9
1175.9
1178.9
1177.4
1176.6
1176.7
1177.4
1177.8
1178.1
1178.0
1175.8
1176.2
1176.2
1176.4
1175.7
1176.3
1176.2
1178.9
1177.4
1179.0
1177.5
1176.0
1178.9
1177.4
1175.9
1179.0
1177.5
1176.0
1177.4
1177.4
1178.9
1177.4
1179.0
1177.5
1176.0
1179.0
1177.4
1175.9
Construction
2" PVC
4" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
"PVC
"PVC
"PVC
"PVC
"PVC
"PVC
"PVC
"PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
2" PVC
1" PVC
1" PVC
3 channel CMT
3 channel CMT
3 channel CMT
3 channel CMT
3 channel CMT
3 channel CMT
3 channel CMT
3 channel CMT
3 channel CMT
1" PVC
1" PVC
1" PVC
1" PVC
3 channel CMT
3 channel CMT
3 channel CMT
3 channel CMT
3 channel CMT
3 channel CMT
-------
Table B1. Method reporting limits for selected analytes during five PRB sampling events and results of selected duplicate analyses.
Technique ICP-OES
Date Al
Well
EPA03
EPA03-DUP
EPA01
EPA01-DUP
TR1
TR1-DUP
EPA06
EPA06-DUP
EPA08
EPA08-DUP
EPA01
EPA01-DUP
TRO
TRO-DUP
T3C
T3C-DUP
7/05
10/05
6/06
9/06
7/07
Sampling Date
7/19/2005
7/19/2005
RPD
10/5/2005
10/5/2005
RPD
6/8/2006
6/8/2006
RPD
6/7/2006
6/7/2006
RPD
9/18/2006
9/18/2006
RPD
9/25/2006
9/25/2006
RPD
7/19/2007
7/19/2007
RPD
7/26/2007
7/26/2007
RPD
0.017
0.017
0.017
0.036
0.030
0.063
0.061
3.2
0.286
0.281
1.8
0.028
0.018
43.5
0.147
0.128
13.8
0.249
0.235
5.8
0.183
0.181
1.1
ND
ND
NA
ND
ND
NA
ICP-OES
Ca
0.017
0.017
0.017
0.024
0.028
39.0
41.0
5.0
52.1
51.9
0.4
73.1
72.4
1.0
42.8
42.7
0.2
51.5
50.1
2.8
38.6
40.0
3.6
3.68
3.73
1.3
14.5
14.4
0.7
ICP-OES
Mg
0.020
0.020
0.020
0.024
0.023
13.8
14.3
3.6
20.8
20.8
0.0
21.9
21.7
0.9
16.7
16.8
0.6
20.3
19.8
2.5
15.3
15.8
3.2
7.29
7.56
3.6
12.4
12.3
0.8
ICP-OES
Mn
0.001
0.001
0.001
0.001
0.002
3.35
3.36
0.3
5.73
5.73
0.0
5.31
5.28
0.6
4.57
4.63
1.3
5.12
5.13
0.2
3.96
4.11
3.7
0.015
0.020
28.6
0.185
0.188
1.6
ICP-OES
Fe
0.006
0.006
0.006
0.005
0.008
4.55
5.71
22.6
6.92
6.90
0.3
27.8
25.2
9.8
8.45
8.43
0.2
6.87
6.77
1.5
4.76
4.88
2.5
ND
ND
NA
ND
ND
NA
ICP-OES
K
0.042
0.042
0.042
0.055
0.092
12.4
12.5
0.8
14.7
14.7
0.0
15.5
15.4
0.6
14.0
13.9
0.7
13.1
13.0
0.8
9.85
9.91
0.6
13.0
13.0
0.0
12.4
12.3
0.8
ICP-OES
Na
0.122
0.122
0.122
0.042
0.091
407
408
0.2
299
299
0.0
263
261
0.8
313
318
1.6
220
227
3.1
153
153
0.0
202
203
0.5
245
242
1.2
ICP-OES
Zn
0.005
0.005
0.005
0.006
0.016
0.80
0.81
1.4
2.96
2.99
1.0
0.021
0.021
0.0
1.19
1.19
0.0
2.28
2.25
1.3
1.93
1.99
3.1
ND
ND
NA
ND
ND
NA
ICP-OES
Si
0.011
0.011
0.011
0.044
0.047
10.0
10.2
2.0
12.2
12.2
0.0
14.3
14.4
0.7
13.0
13.0
0.0
11.5
11.5
0.0
9.03
9.29
2.8
0.44
0.43
1.8
2.12
2.16
1.9
ICP-OES
As
0.004
0.004
0.004
0.005
0.008
38.9
38.8
0.3
28.4
28.4
0.0
18.1
18.2
0.6
34.0
34.1
0.3
27.3
27.5
0.7
19.3
20.2
4.6
0.134
0.133
0.7
0.724
0.739
2.1
IC-ICP-MS
As(III)
0.014
0.014
0.013
0.015
0.015
34.4
34.8
1.2
23.7
22.3
6.1
3.16
3.57
12.2
22.0
22.2
0.9
19.3
18.1
6.4
15.5
15.7
1.3
0.141
0.117
18.6
1.53
1.54
0.7
IC-ICP-MS
As(V)
0.014
0.014
0.013
0.015
0.015
0.97
0.96
1.0
4.19
4.32
3.1
10.5
10.7
1.9
8.77
7.25
19.0
5.60
5.17
8.0
4.08
4.40
7.5
ND
ND
NA
ND
ND
NA
CE
Cl
0.100
0.100
0.100
0.100
0.113
68.6
67.7
1.3
65.7
63.5
3.4
42.1
42.5
0.9
42.6
41.5
2.6
36.1
36.9
2.2
41.0
33.3
20.7
32.7
33.2
1.5
34.8
33.4
4.1
CE
S°4
0.100
0.100
0.100
0.100
0.137
645
636
1.4
621
619
0.3
470
476
1.3
628
623
0.8
547
554
1.3
667
536
21.8
205
206
0.5
223
213
4.6
CD
—5
_J
Q_
— •
X
00
RPD is the relative percent difference used to compare duplicate results, RPD = (|C2 - Cl |)/((C1 + C2)/2) x 100, where Cl is the concentration in the sample and C2 is the
concentration in the duplicate. The replicate results are in good agreement. Less than 5% difference is observed in 90 out 104 comparisons; better than 10% difference is
observed in 95 out of 104 observations. As expected, larger percent differences are observed in measurements approaching the detection limit and are sometimes associated
with the arsenic speciation analysis.
-------
Table B2. Results of blank tests and pump rinsate tests.
Field Blank
Field Blank
Field Blank
Field Blank
Field Blank
Pump Rinse
PBTW-1
Pump Rinse
EPA05
Technique
Date
7/05
10/05
6/06
9/06
7/07
8/02
8/02
ICP-OES
Al
ND
ND
ND
ND
ND
ND
0.15
ICP-OES
Ca
ND
ND
0.26
ND
ND
0.88
0.44
ICP-OES
Mg
ND
ND
0.05
ND
ND
0.45
0.17
ICP-OES
Mn
ND
ND
ND
ND
ND
0.02
0.02
ICP-OES
Fe
0.02
ND
ND
ND
ND
0.33
0.71
ICP-OES
K
0.06
ND
ND
ND
ND
ND
ND
ICP-OES
Na
ND
0.18
0.70
ND
ND
1.85
0.57
ICP-OES
Zn
ND
ND
ND
ND
ND
ND
ND
ICP-OES
Si
ND
0.55
0.54
ND
ND
0.26
0.35
ICP-OES
As
ND
0.007
ND
0.005
ND
ND
ND
IC-ICP-MS
As(III)
ND
ND
ND
ND
ND
NA
NA
IC-ICP-MS
As(V)
ND
ND
ND
ND
ND
NA
NA
CE
Cl
ND
0.20
ND
ND
ND
NA
NA
CE
so4
ND
ND
0.67
ND
ND
NA
NA
ND, not detected. NA, not analyzed. All results in mg/L.
The majority of the field based analyses indicate ND or near MDL values. The analytes with detectable values are likely reflective of trace components in distilled water purchased
locally for field decontamination. Low levels of arsenic in blanks from 10/2005 and 9/2006 are likely due to artifacts after running samples with high arsenic concentrations (>50
mg/L). The Fultz pump was rinsed in between wells with 1 gallon of distilled water. At the end of the rinse cycle, water was collected and analyzed in wells PBTW-1 and EPA05 to
evaluate decontamination effectiveness.
-------
Appendix C
Table Cl. Thermodynamic data for arsenic oxyanions.
Reaction
As(III) species
H3AsO3(aq) = H2AsO3' + H+
H2AsO3- = HAsO32' + H+
HAsO32' = AsO33' + H+
As(V) species
H3AsO4(aq) = H2AsO; + H+
H2AsO; = HAsO/- + H+
HAsO42-=AsO43- + H+
Redox
H3As04(aq) + H2(g) =
H3As03(aq) + H20(l)
MINTEQA21
-9.38
-14.35
ND
-2.24
-7.01
-11.94
22.96
WATEQ4P
-9.36
-14.70
-15.70
-2.25
-7.18
-11.79
23.40
LLNL3
-9.23
-11.01
ND
-2.25
-6.75
-11.90
22.56
Nordstrom
and Archer
(2003)
-9.15
-14.70
-15.70
-2.30
-7.16
-11.65
19.35
Raposo
et al. (2004)
-2.25
-7.06
-11.58
Zakaznova-
Herzog et al.
(2006)
-9.25
1 Allison, J. D., Brown, D. S., and Nova-Gradac, K. J. (1990). MINTEQA2/PRODEFA2, A geochemical assessment model for
environmental systems. USEPA, EPA/600/3-91/021.
2 Ball, J. W. and Nordstrom, D. K. (1991). User's manual for WATEQ4F, with revised thermodynamic data base. USGS Open File
Report 91-183.
3 Dataset from Lawrence Livermore National Laboratory, thermo.dat. Wolery, T. J. (1992). EQ3/6, a software package for
geochemical modeling of aqueous systems, package overview and installation guide (version 7.0). Technical Report UCRL-
MA-110662, Part 1, Lawrence Livermore National Laboratory.
Nordstrom, D. K. and Archer, D. G. (2003). Arsenic thermodynamic data and environmental geochemistry. In Arsenic in Ground
Water: Geochemistry and Occurrence, Welch, A. H. and Stollenwerk, K. G., eds. Kluwer Academic Publishers, The Netherlands.
Raposo, J. C., Zuloaga, O., Olazabal, M. A., and Madariaga, J. M. (2004). Study of the precipitation equilibria of arsenate anion
with calcium and magnesium in sodium perchlorate at 25°C. Applied Geochemistry, v. 19, p. 855-862.
Zakaznova-Herzog, V. P., Seward, T. M., and Suleimenov, O. M. (2006). Arsenous acid ionization in aqueous solutions from 25 to
300 °C. GeochimicaetCosmochimicaActa, v. 70, p. 1928-1938.
-------
Appendix D
Ground-Water Elevation Hydrographs
S 11 SO 30
ra
° 117950
2002 2003 20C4 2COC 2006 2007 200
Date
Figure D1. Hydrograph of ground-water elevations
measured at well DH-17.
<
E
TO
C
O
006 2007 2003
Date
Figure D3. Hydrograph of ground-water elevations
measured at well EPA06.
<
E
CO
s
<
_
LU
fc
117950
2:06 zo:
Date
C06 20"
Date
Figure D2. Hydrograph of ground-water elevations
measured at well EPA02.
Figure D4. Hydrograph of ground-water elevations
measured at well PBTW-2.
-------
O
(5 117900
Date
Figure D5. Hydrograph of ground-water elevations
measured at well TR8.
-------
Appendix E
Potentiometric Surface in Vicinity of PRB
262550-
0)
.E 262450-
I
262350-
414300
414400
414500
Easting (m)
Figure El. Potentiometric surface at the water table in vicinity of PRB on June 19, 2002.
' }J*-T(jf
262550-
O)
.E 262450H
262350-
414300 414400 414500
Easting (m)
Figure E2. Potentiometric surface at the water table in vicinity of PRB on September 26, 2002.
-------
262550-
O)
.E 262450-
.C
t:
o
262350-
414300
414400
414500
Easting (m)
Figure E3. Potentiometric surface at the water table in vicinity of PRB on August 14, 2003.
262550
414300 414400 414500
Easting (m)
Figure E4. Potentiometric surface at the water table in vicinity of PRB onMay 31, 2005.
-------
414300 414400 414500
Easting (m)
Figure E5. Potentiometric surface at the water table in vicinity of PRB on July 20, 2005.
262550-
O)
.E 262450
•e
o
262350-
414300
414400
414500
Easting (m)
Figure E6. Potentiometric surface at the water table in vicinity of PRB on October 6, 2005.
-------
262550
O)
.£ 262450-
.C
tr
o
262350-
414300 414400 414500
Easting (m)
Figure El. Potentiometric surface at the water table in vicinity of PRB on June 6, 2006.
262550-
D)
.E 262450
.C
-c
O
262350-
414300 414400 414500
Easting (m)
Figure E8. Potentiometric surface at the water table in vicinity of PRB on September 18, 2006.
-------
262550
05
.E 262450-
€
o
262350-
414300 414400 414500
Easting (m)
Figure E9. Potentiometric surface at the water table in vicinity of PRB on January 24, 2007.
262550-
O)
.£ 262450
.C
tr
o
262350-
414300 414400 414500
Easting (m)
Figure E10. Potentiometric surface at the water table in vicinity of PRB on July 18/19, 2007.
-------
262500
CD
.E 262450
262400-
414400
I
414450
Easting (m)
414500
Figure Ell. Potentiometric surface at the water table in vicinity of PRB on October 1, 2007.
262550-
O)
.£ 262450-
tr
O
262350-
414300
414400
414500
Easting (m)
Figure E12. Potentiometric surface at the water table in vicinity of PRB on April 1, 2008.
-------
262550-
D3
.E 262450
.C
tr
o
262350-
414300 414400 414500
Easting (m)
Figure E13. Potentiometric surface at the water table in vicinity of PRB on June 24, 2008.
-------
-------
Recycled/Recyclable
Printed with vegetable-based ink on
paper that contains a minimum of
50% post-consumer fiber content
processed chlorine free
------- |