EPA 600/R-08/093 | September 2008 | www.epa.gov/ada

Field Application of a Permeable 

Reactive Barrier for Treatment of 

Arsenic in Ground Water

Office of Research and Development

National Risk Management Research Laboratory, Ada, Oklahoma 74820

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 environmentally 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. 

ii 
 

Robert W. Puls, Acting Director 

Ground Water and Ecosystems Restoration Division 

National Risk Management Research Laboratory 

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 costeffectiveness 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. 

iii 

Contents Contents
 

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 zerovalent 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
 

v 
Figures
 

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 

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. ........................................................... 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 flowmeter test 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 As2

O5 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
 

vi 
Figure 19. Comparison of Eh values of ground water measured using a platinum 

electrode and calculated using the As(III)-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 EPA10 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
 

vii 
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 D1. 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 E1. 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 water table in vicinity of PRB on August 14, 2003. ........ 60
 

Figure E4. Potentiometric surface at the water table in vicinity of PRB on May 31, 2005. .......... 60
 

Figure E5. Potentiometric surface at the water table in vicinity of PRB on July 20, 2005. .......... 61
 

Figure E6. Potentiometric surface at the water table in vicinity of PRB on October 6, 2005. ........ 61
 

Figure E7. Potentiometric surface at the water table 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 water table in vicinity of PRB on July 18/19, 2007. ....... 63
 

Figure E11. Potentiometric surface at the water table in vicinity of PRB on October 1, 2007. ....... 64
 

Figure E12. Potentiometric surface at the water table in vicinity of PRB on April 1, 2008. ......... 64
 

Figure E13. Potentiometric surface at the water table in vicinity of PRB on June 24, 2008. ......... 65
 

viii 
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 A1. Wells used in the design and assessment of the PRB. ............................... 52
 

Table B1. 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 C1. Thermodynamic data for arsenic oxyanions. ..................................... 56
 

ix 
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. DNDCAT 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 Köber (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. 

x 
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 L-1 in wells located hydraulically upgradient of the PRB. 

Within the PRB, arsenic concentrations are reduced to 2 to <0.01 mg L-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. 

xi 

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 fluidflow 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 

1.0 

Introduction
 

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, pilotscale field testing, and full-scale field applications.

1 
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, 

H AsO -

, HAsO 2-, and AsO 3-. In anoxic waters, the 2 4 4 4 

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 μM, 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, 

Magalhães (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 ironreducing 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 mineralwater 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 L-1, decreased from the 

previous level of 0.05 mg L-1. This revision of the MCL 

recognizes the detrimental health effects associated with 

arsenic in drinking water, including bladder, skin, and

2 
lung cancers, diabetes, and neurological dysfunction 

(National Research Council, 1999). Elevated 

concentrations of arsenic from natural sources (>0.05 

mg L-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 L-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; Köber 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, 2001b). 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, 2001a; 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, coprecipitation, 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; Köber et al., 2005). This pilot-scale study allows 

for an examination of arsenic uptake processes in a 

complex field setting. 

3 
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. 

4 
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. 

5 
4.0 

PRB Installation 

The pilot-PRB is a 9.1 m 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 

6.1 m 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 ~1-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 

6 
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 

7 
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 7.6 m and 14.6 m 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 

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.
	

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 A1 (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. Groundwater 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, oxidationreduction 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, n=132) as determined using a Hach 

turbidimeter (Model 2100P). Field measurements were 

made for sulfide and ferrous iron using the methylene 

blue and 1,10-phenanthroline colorimetric methods, 

respectively (Hach DR/2010). Alkalinity measurements 

9 
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 greenmethyl red endpoint. 

Filtered samples (0.45 μm, 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 HCl, and retained 

in amber-plastic bottles. Speciation analysis was carried 

out using ion chromatography (IC) 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 L-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 in Eary (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). 

10 
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 GeoprobeTM. 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 subsamples 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-μm filter paper (polycarbonate). The 

separated particles were mounted on a zero-background 

quartz plate and scanned with Fe Kα 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 semiquantitative 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 

11 
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 DNDCAT, 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 I1 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 PNCCAT 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 Sample Type Beam 

Line XAS Data 

30 ft Boring PNC XANES 

35 ft Boring PNC XANES 

45 ft Boring PNC XANES 

50 ft Boring PNC XANES 

Core 1 30-34 ft fines PRB Core DND XANES 

Core 1 44-48 ft PRB Core DND XANES 

Core 2 30-34 ft PRB Core DND XANES 

Core 2 38-42 ft fines PRB Core DND XANES 

Core 2 38-45 ft fines PRB Core DND XANES 

Core 3 30-38 ft PRB Core DND XANES 

Core 3 38-42 ft PRB Core DND XANES 

Core 3 38-42 ft fines PRB Core DND XANES 

Core 4 30-34 ft PRB Core DND XANES 

APBH 4-2 10-11.5 ft Source Zone DND XANES 

APBH 4-1 5-6.5 ft Source Zone DND XANES 

TW 1-5 20-25 ft Source Zone DND XANES 

TW 1-8 38-40 ft Source Zone DND XANES 

TW 1-6 25-30 ft Source Zone DND XANES 

Arsenopyrite Reference DND XANES 

As2

O3 Reference DND XANES 

Arsenian Pyrite Reference DND XANES 

As(III) sorbed to

Ferrihydrite Reference DND XANES 

As(V) sorbed to

Ferrihydrite Reference DND XANES 

Elemental As Reference DND XANES 

As(III) sorbed FeS Reference PNC 

As(III) sorbed FeS

Oxidized Reference DND XANES 

As(III) sorbed

to FeS Oxidized Reference PNC XANES 

(duplicate) 

As(III) sorbed to

Pyrite Reference DND XANES 

Scorodite 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). 

12 
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 Edge Position 

(eV) 

White Line 

(eV) 

Elemental As 

Orpiment 

As(III) sorbed FeS 

11866.7 

11868.7 

11867.6 

11868.7 

11870.1 

11870.5 

As(III) sorbed to 

Ferrihydrite 

Sodium Arsenite

11869.9 

 11870.1 

11871.9 

11871.8 

Enargite 

Sodium Arsenate 

As(V) sorbed to 

Ferrihydrite 

11870.2 

11874.0 

11873.8 

11871.7 

11875.5 

11875.5 

Scorodite 11874.0 11875.8 

Notes: Edge position is defined as the maximum in the firstderivative of the energy vs. absorption function. White 

line position is defined as the maximum in the energy vs. 

absorption function. 

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. 

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). 

Measurement 

Date 

Multi-Point Estimate Three-Point Estimate 

Direction Direction Magnitude Magnitude (degrees) (degrees) 

10/6/05 0.005 336 0.005 335 

6/6/06 0.008 342 0.008 323 

9/18/06 0.005 338 0.005 339 

7/19/07 0.005 331 0.005 345 

10/1/07 0.005 335 0.005 335 

4/1/08 0.006 329 0.006 331 

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 

14 

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). 

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 min-1 to 40 L min-1. 

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 min-1 to 

7 L min-1. 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 

15 
K (∆Q − ∆q )∆z i i i i = ; i = 1, 2... n. K QP b 

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, Ki is the horizontal hydraulic 

conductivity of interval i, ∆Qi is induced flow from 

interval i, ∆qi is ambient flow from interval i, ∆zi is the 

thickness of interval i, 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. 

16 
 

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 L-1) and arsenic 

(>100 mg L-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 L-1), and arsenic concentrations 

(<50 mg L-1); sulfate becomes the dominant anion and 

the proportion increases of calcium and magnesium 

relative to sodium (Figure 9). 

6.0 

Results and Discussion
 

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+

-SO4

2- type, with Na+

 (150-900 mg L-1) 

and SO4

2- (300-1000 mg L-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 

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 
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 L-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 11c,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 11c), a transition in arsenic concentration 

is observed at a depth of about 12 m 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 L-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. 

18 

Figure 10. Long-term trends in the major ion chemistry of ground water collected from monitoring well PBTW-1.
	
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 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). 

20 
 

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 L-1), arsenic is generally present at lower 

concentrations (<10 mg L-1) as arsenate. 

Figure 13. Arsenic concentrations in solutions saturated 

with As2O5 and As2O3 and site ground water 

(open red circles) compared to the MCL for 

arsenic. 

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. 

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 L-1 

and 40 mg L-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 L-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 L-1, which agrees reasonably with the bulk well 

concentration noted above (~28.5 mg L-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 nearneutral in pH (6.1-6.7), anoxic/suboxic (generally 

<0.5 mg DO L-1), moderately reducing (Eh values from 

130 to 200 mV), and contains moderate concentrations 

of ferrous iron (<10 mg L-1) and dissolved organic 

carbon (<5 mg L-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 L-1. These high concentrations 

22 
Table 4. Ground-water parameters in selected wells.

Parameter Date Al (mg/L) Ca (mg/L) Mg (mg/L) Mn (mg/L) Fe (mg/L) K(mg/L) Na (mg/L) Zn (mg/L) Si (mg/L) Cl (mg/L) SO4 (mg/L) HCO3 (mg/L) T(°C) pH(SU) Eh(mV) DO(mg/L) 

Source

DH-33 Jun-06 0.09 2.96 0.836 0.10 0.27 6.41 786 0.06 3.10 60.0 1050 672 11.7 10.2 157 0.3 

DH-33 Jul-07 0.07 10.6 8.43 0.31 0.36 5.85 869 <0.02 3.17 74.8 1050 917 11.7 9.39 15 0.5 DH-21 Jul-07 0.07 14.2 6.82 0.62 0.06 14.5 807 0.03 6.30 25.4 557 1383 10.9 8.81 189 3.6 DH-34 Jul-07 0.05 4.13 1.78 0.02 0.12 19.0 548 <0.02 6.11 30.2 502 786 11.1 9.58 68 0.6 

Upgradient

PBTW-1 Jun-06 <0.05 32.5 9.45 1.4 0.40 11.9 509 0.10 7.8 67.3 907 420 13.4 7.50 173 0.2 

EPA08 Jun-06 0.23 45.2 18.3 4.7 5.5 13.9 249 2.1 13.5 42.5 514 300 13.6 6.50 207 0.1 EPA08 Sep-06 0.25 51.1 20.3 5.1 6.9 13.1 220 2.3 11.5 37.2 545 221 13.1 6.20 159 <0.1 EPA08, 33ft Jul-07 1.30 72.1 28.4 9.4 8.2 15.1 223 7.9 16.6 34.0 593 180 13.2 5.85 210 1.0 EPA08, 37ft Jul-07 0.28 59.1 23.4 6.2 6.9 14.4 222 3.5 14.0 33.5 510 196 13.2 5.89 221 0.5 EPA08, 42ft Jul-07 0.22 55.9 22.5 5.8 5.6 14.5 228 3.3 13.7 34.8 509 206 13.2 6.25 224 0.4 

PRB 

T3A Jul-05 0.03 26.0 7.94 0.99 3.52 3.4 126 0.03 0.26 38.8 63 325 nm 7.42 -382 <0.1 

T3A, early Jun-06 0.07 4.97 1.55 0.04 0.50 9.5 255 <0.02 1.29 50.3 81 385 13.9 10.10 42 <0.1 T3A Sep-06 <0.05 10.8 12.6 0.07 <0.01 10.5 281 <0.02 1.99 42.6 365 372 13.2 9.34 -43 <0.1 T3A, late Jul-07 <0.05 3.15 3.73 0.01 <0.01 15.1 233 <0.02 0.54 34.6 192 366 16.6 10.07 181 0.3 S5 Jun-06 0.146 4.63 0.61 0.03 0.28 4.59 95 0.03 3.30 23.0 52 175 13.5 9.10 113 <0.1 S8 Jun-06 0.113 3.22 4.95 0.01 0.22 7.3 120 <0.02 1.28 19.7 191 106 14.2 6.80 25 <0.1 S2 Jul-07 <0.05 1.25 <0.02 0.00 0.02 11.1 194 <0.02 0.35 23.8 8.4 368 15.9 10.31 -134 0.4 S4 Jul-07 <0.05 14.5 0.31 0.01 0.02 17.6 155 <0.02 0.20 17.5 9.7 337 14.0 9.92 -75 0.5 

23 
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, 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 pilotPRB was difficult to filter (using 0.45 μm disc filters) 

and slightly malodorous. After 12 months, however, 

DOC concentrations decreased to near-background 

levels (mean value 9 mg L-1), and after 15 months 

DOC concentrations in the PRB were indistinguishable 

from concentrations present upgradient of the PRB 

(mean value 5 mg L-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 L-1. Within the PRB, arsenic concentrations 

have generally been <0.50 mg L-1. Of 80 samples 

collected from the pilot-PRB, only 11 samples exceeded 

0.50 mg As L-1 (see Figure 17b); 62 samples had 

concentrations of arsenic at or below 0.50 mg L-1; and, 

24 samples were at or below the MCL for arsenic of 

0.01 mg L-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) + 2H2

O = 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

-

, HAsO4

2-, 

and H3AsO3

0

. 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. 

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 ΣAs=10-4. 

25 
H AsO 0

 + 2H+ + 2e- = H AsO 0 + H O. 3 4 3 3 2

 

 

Oxidation-reduction potentials were calculated from the 

measured arsenic speciation based on the following halfcell reaction: 

The activities of H AsO 0 and H AsO 0 were calculated 3 4 3 3 

based on the measured arsenic speciation, pH, and 

calculated ionic strength. The equilibrium expression 

was solved for ae- 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 Rüde 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 oxidationreduction potential. 

Figure 19. Comparison of Eh values of ground water 

measured using a platinum electrode and 

calculated using the As(III)-As(V) redox 

couple. 

Monitoring well EPA09 is located immediately 

downgradient of the pilot-PRB. This well was sampled 

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 L-1, and 15.5 mg L-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 depthresolved 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. 

26 
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). 

Figure 21. Ground-water elevations obtained using data from a pressure transducer/data logger installed in well DH17 located approximately 30 m upgradient of the PRB. 

27 
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 groundwater 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. 

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. 

28 
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. 

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 
 

The downgradient transport of the bromide was 

monitored using nine observation wells (TR3, TR4-2, 

TR5-2, TR6-2, TR7, TR8, TR10, 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. 

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. 

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 TR10 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. 

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 m d-1 to 41.1 m d-1 with an average of 33.2 m d-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 intervalspecific hydraulic conductivity of these materials from 

the borehole flowmeter surveys. 

30 
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. 

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 October 

2005 

September 

2006 July 2007 

EPA10 49 

S01 88 

S02 75 

S03 98 98 

S04 7.3 11 

S05 56 42 

S06 100 110 

S07 95 92 

S08 70 69 

TR1 32 43 

TR2 55 

TR3 51 

TR7 43 

TR8 47 

TR9 2.1 

TR10 34 

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. 

31 
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. 

32 

Figure 29. Results of pneumatic slug tests within the PRB after 5 months and 16 months.
	
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. 

Figure 31. Hydraulic conductivity distribution estimated 

for materials adjacent to the screened 

intervals of wells EPA02, EPA08, EPA09, 

and EPA10 based on characterization using 

borehole flowmeter techniques. 

The borehole flowmeter technique was also applied to 

characterization of materials adjacent to well EPA10 

(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 EPA10 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. 

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. 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 depthdependent volumetric flow velocity (Darcy velocity) by 

depth-dependent arsenic concentrations in wells EPA08 

and EPA09 (Figure 33): 

Arsenic Flux = (Ki ·dh/dx) C, 

where Ki is the interval specific hydraulic conductivity, 

dh/dx 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 m-2d-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-1. 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 

1177 m AMSL. 

34 
 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 ferriciron 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 L-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). 

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. 

Elevated concentrations of sulfate (~530 mg L-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 L-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). 

Figure 36. Sulfate removal within the PRB as a function 

of time and depth. Sulfate concentration 

entering the PRB, C0, is 530 mg L-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 Köber 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 L-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 L-1 and 0.42 mg L-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). 

36 
 

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 solidphase 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 (Köber 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 μm; green 

circles, Fe=1 μm). 

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 μg L-1), Cd 

(50 μg L-1), Co (20 μg L-1), and Ni (10 μg L-1). 

Concentrations of Cd, Co, and Ni within ground water 

from the PRB are below detection (generally < 1 μg L-1), 

and Zn concentrations have decreased to <50 μg 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. 

37 
 

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 μg L-1. Selenium concentrations 

within the PRB span from about 5 μg L-1 to 18 μg L-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. 

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 41a 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 10× 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. 

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 
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. 

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 

for As(V)-O. 

40 
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 sulfate-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 

41 
of As(III)-O and As(V)-O. Figure 44 shows the depthdependent 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, sulfidebearing minerals are not concentrated enough to be 

detectable by bulk XRD methods. 

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. 

Figure 44. a) Depth-dependent concentrations of total 

42 

Figure 45. Powder x-ray diffraction scans of aquifer materials collected from a well boring taken adjacent to the pilotPRB. 
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. 

Core 

Coordinated fitting model 

Depth (ft) As(V)-O As(III)-O As(III)-S R-Factor 

coordinated 

Sorbed fitting model 

As(V)-Fh As(III)-Fh As(III)-FeS R-Factor 

sorbed 

PRB1 

PRB1 

PRB2 

PRB2 

PRB2 

PRB3 

PRB3 

PRB3 

PRB4 

30-34F 

44-48 

30-34B 

38-42F 

38-45F 

30-38A 

38-42 

38-42F 

30.34B 

28.8 

18.5 

35.9 

14.4 

22.4 

37.3 

26.1 

20.1 

39.9 

41.7 

55.1 

39.3 

60.5 

60.6 

37.3 

48.1 

29.1 

45.8 

29.4 

26.4 

24.8 

25.1 

17.1 

25.4 

25.8 

50.8 

14.3 

0.004 

0.015 

0.006 

0.011 

0.011 

0.019 

0.017 

0.011 

0.005 

31.7 

19.3 

40.6 

19.1 

26.6 

41.8 

29.5 

16.6 

48.4 

30.4 

59.3 

31.9 

54.1 

62.2 

41.8 

35.9 

18.7 

26.2 

37.9

21.4 

27.5 

26.9 

11.2 

16.4 

34.6 

64.7 

25.3 

 0.0010 

0.0013 

0.0015 

0.0008 

0.0008 

0.0060 

0.0037 

0.0013 

0.0017 

APBH 4-1 

APBH 4-2 

TW 1-5 

TW 1-6 

TW 1-8 

1.5- 6.5 

10- 11.5 

20- 25 

25- 30 

38- 40 

93.5 

72.0 

50.0 

65.8 

73.4 

6.5 

28.0 

27.0 

34.2 

26.6 

0.0 

0.0 

23.0 

0.0 

0.0 

0.015 

0.015 

0.011 

0.015 

0.011 

99.4 

79.8 

57.0 

75.4 

81.8 

0.6 

20.2 

43.0 

24.6 

18.2 

0.0 

0.0 

0.0 

0.0 

0.0 

0.0060 

0.0050 

0.0015 

0.0050 

0.0030 

Boring 

Boring 

Boring 

Boring 

30 

35 

45 

50 

28.7 

30.2 

50.3 

51.1 

0.9 

8.5 

49.7 

48.9 

70.4 

61.2 

0.0 

0.0 

0.009 

0.011 

0.014 

0.016 

18.0 

22.7 

63.9 

64.1 

0.0 

0.0 

28.7 

35.4 

82.0 

77.3 

7.4 

0.5 

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. 

43 
Figure 46. Comparison of LCF fitting results using a) 

coordinated or b) sorbed reference materials 

for PRB 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. 

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. 
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. 

45 
 

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 L-1 in wells located 

hydraulically upgradient of the PRB. Within the PRB, 

arsenic concentrations are reduced to 2 to <0.01 mg L-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 L-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 L-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 d-1, 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. Solidphase 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 

46 
precipitation is an undesirable consequence of 

zerovalent iron applications and is expected to reduce 

long-term reactive and hydraulic performance. 

Microbial reduction of sulfate to sulfide and 

precipitation of low-density iron sulfides within the 

reactive medium creates additional mineral surfaces 

for arsenic removal. 

• Source control measures will reduce arsenic loading to 

the PRB and implementation of such measures should 

result in increased PRB lifetimes. The increase in 

effective lifetime should be directly proportional to the 

level of concentration reduction achieved by isolating 

and/or treating the source of contamination. 

47 
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 unconfined 

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 in formations 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 As2

S3 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. 

48 
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. 

Soil 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 Fe0 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). 

Köber, 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 83: 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 groundwater resources of the Helena Valley, Montana. Circular 

83, U.S. Geological Survey, Reston, VA, 1951. 

Magalhães, 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). 

McArthur, 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). 

49 
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. Solidphases 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. 

Saha, 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. Batchmixed 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). 

Rüde, T. R. and S. Wohnlich. “The couple of As(V) – 

As(III) as a redox indicator.” In Redox: fundamentals, 

processes, and applications, J. Schüring 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 

(2001a). 

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 (2001b). 

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 B129: 297-303 (2006). 

USEPA. Technology alternatives for the remediation 

of soils contaminated with As, Cd, Cr, Hg, and Pb, 

EPA/540/S-97/500. U. S. Environmental Protection 

Agency, 1997. 

USEPA. The QTRACER2 program for tracerbreakthrough curve analysis for tracer tests in karstic 

and other hydrologic systems, EPA/600/R-02/001. 

50 
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/R03/045a. 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, P.A. 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. Wallschläger, 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 zerovalent 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 

borehole flowmeter, 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). 

51 
 

Appendix A 

Table A1. Wells used in the design and assessment of the PRB. 

Existing Wells: 

WELL 

Ground Surface 

Elevation 

(m AMSL) 

Top of Screen 

(m BLS) 

Bottom of Screen 

(m BLS) 

Top of Screen 

(m AMSL) 

Bottom of Screen 

(m AMSL) Construction 

DH-8 1192.61 11.9 14.9 1180.7 1177.7 4” PVC 

DH-17 1189.34 9.4 12.5 1179.9 1176.8 4” PVC 

DH-21 1190.53 5.8 8.8 1184.7 1181.7 4” PVC 

DH-24 1187.91 8.2 10.7 1179.7 1177.2 4” PVC 

DH-27 1191.53 5.8 8.8 1185.7 1182.7 4” PVC 

DH-33 1191.43 6.1 9.1 1185.3 1182.3 2” PVC 

DH-34 1190.75 6.1 9.1 1184.7 1181.6 2” PVC 

DH-36 1189.44 6.4 9.4 1183.0 1180.0 2” PVC 

DH-49 1188.15 7.3 10.4 1180.8 1177.8 2” PVC 

DH-50 1188.35 7.3 10.4 1181.0 1178.0 2” PVC 

DH-51 1188.36 7.3 10.4 1181.0 1178.0 2” PVC 

DH-56 1204.58 21.3 25.9 1183.2 1178.7 2” PVC 

DH-63 1188.64 7.3 11.9 1181.3 1176.8 2” PVC 

DH-64 1188.20 13.7 16.8 1174.5 1171.4 2” PVC 

DH-66 1191.26 11.6 14.6 1179.7 1176.6 2” PVC 

STW-1 1188.65 10.1 11.6 1178.6 1177.1 2” PVC 

STW-4 1188.14 9.8 11.3 1178.4 1176.9 2” PVC 

STW-7 1188.67 7.6 12.2 1181.0 1176.5 2” PVC 

STW-9 1188.23 10.7 12.2 1177.6 1176.0 2” PVC 

PRB1 1190.24 10.7 15.2 1179.6 1175.0 2” PVC 

PRB2 1188.80 11.3 15.8 1177.5 1172.9 2” PVC 

PRB3 1190.96 11.0 15.5 1180.0 1175.4 2” PVC 

Note: AMSL = Above Mean Sea Level. 

BLS = Below Land Surface. 

52 
 

Wells Installed by USEPA: 

WELL 

Ground Surface 

Elevation 

(m AMSL) 

Top of 

Screen 

(m BLS) 

Bottom 

of Screen 

(m BLS) 

Top of 

Screen 

(m AMSL) 

Bottom 

of Screen 

(m AMSL) 

Construction 

EPA01 1189.72 10.1 14.6 1179.7 1175.1 2” PVC 

EPA02 1189.59 10.1 14.6 1179.5 1175.0 4” PVC 

EPA03 1189.35 10.1 14.6 1179.3 1174.7 2” PVC 

EPA04 1189.41 10.1 14.6 1179.4 1174.8 2” PVC 

EPA05 1188.89 10.1 14.6 1178.8 1174.3 2” PVC 

EPA06 1189.45 10.1 14.6 1179.4 1174.8 2” PVC 

EPA07 1188.99 10.1 14.6 1178.9 1174.4 2” PVC 

EPA08 1189.63 10.1 14.6 1179.6 1175.0 2” PVC 

EPA09 1189.60 10.1 14.6 1179.5 1175.0 2” PVC 

EPA10 1189.57 9.1 14.3 1180.4 1175.2 2” PVC 

PBTW-1 1189.29 9.4 13.7 1179.8 1175.6 2” PVC 

PBTW-2 1189.01 9.8 16.5 1179.3 1172.6 2” PVC 

S01 1189.57 13.0 13.7 1176.6 1175.9 1” PVC 

S02 1189.57 9.9 10.7 1179.7 1178.9 1” PVC 

S03 1189.58 13.0 13.7 1176.6 1175.9 1” PVC 

S04 1189.57 9.9 10.7 1179.7 1178.9 1” PVC 

S05 1189.59 11.4 12.2 1178.2 1177.4 1” PVC 

S06 1189.57 12.2 13.0 1177.4 1176.6 1” PVC 

S07 1189.61 12.2 13.0 1177.4 1176.7 1” PVC 

S08 1189.61 11.4 12.2 1178.2 1177.4 1” PVC 

T1A 1189.70 9.6 11.9 1180.1 1177.8 2” PVC 

T1B 1189.72 9.4 11.6 1180.4 1178.1 2” PVC 

T1C 1189.72 9.4 11.7 1180.3 1178.0 2” PVC 

T2A 1189.58 9.2 13.8 1180.4 1175.8 2” PVC 

T2B 1189.57 8.8 13.4 1180.8 1176.2 2” PVC 

T2C 1189.59 8.8 13.4 1180.8 1176.2 2” PVC 

T3A 1189.56 8.6 13.2 1181.0 1176.4 2” PVC 

T3B 1189.55 9.3 13.8 1180.3 1175.7 2” PVC 

T3C 1189.56 8.7 13.2 1180.9 1176.3 2” PVC 

TR1 1189.61 10.4 13.4 1179.2 1176.2 2” PVC 

TR2 1189.56 9.9 10.7 1179.7 1178.9 1” PVC 

TR3 1189.58 11.4 12.2 1178.1 1177.4 1” PVC 

TR4-1 1189.57 10.0 10.6 1179.6 1179.0 3 channel CMT 

TR4-2 1189.57 11.5 12.1 1178.1 1177.5 3 channel CMT 

TR4-3 1189.57 13.0 13.6 1176.6 1176.0 3 channel CMT 

TR5-1 1189.57 10.0 10.6 1179.6 1178.9 3 channel CMT 

TR5-2 1189.57 11.5 12.1 1178.0 1177.4 3 channel CMT 

TR5-3 1189.57 13.1 13.7 1176.5 1175.9 3 channel CMT 

TR6-1 1189.57 10.0 10.6 1179.6 1179.0 3 channel CMT 

TR6-2 1189.57 11.5 12.1 1178.1 1177.5 3 channel CMT 

TR6-3 1189.57 13.0 13.6 1176.6 1176.0 3 channel CMT 

TR7 1189.59 11.4 12.2 1178.2 1177.4 1” PVC 

TR8 1189.60 11.4 12.2 1178.2 1177.4 1” PVC 

TR9 1189.60 9.9 10.7 1179.7 1178.9 1” PVC 

TR10 1189.61 11.4 12.2 1178.2 1177.4 1” PVC 

TR11-1 1189.60 10.0 10.6 1179.6 1179.0 3 channel CMT 

TR11-2 1189.60 11.5 12.1 1178.1 1177.5 3 channel CMT 

TR11-3 1189.60 13.0 13.6 1176.6 1176.0 3 channel CMT 

TR12-1 1189.60 10.0 10.6 1179.6 1179.0 3 channel CMT 

TR12-2 1189.60 11.6 12.2 1178.1 1177.4 3 channel CMT 

TR12-3 1189.60 13.1 13.7 1176.5 1175.9 3 channel CMT 

53 
Technique

Date 

ICP-OES ICP-OES ICP-OES ICP-OES

Al Ca Mg Mn 

ICP-OES

Fe 

ICP-OES

K 

ICP-OES

Na 

ICP-OES

Zn 

ICP-OES

Si 

ICP-OES

As 

IC-ICP-MS

As(III) 

IC-ICP-

As(V

7/05 

10/05 

6/06 9/06 

7/07 

0.017 0.017 0.020 0.001 0.017 0.017 0.020 0.001 

0.017 0.017 0.020 0.001 0.036 0.024 0.024 0.001 

0.030 0.028 0.023 0.002 

0.006 0.006 0.006 0.005 

0.008 

0.042 0.042 0.042 0.055 

0.092 

0.122 0.122 0.122 0.042 

0.091 

0.005 0.005 0.005 0.006 

0.016 

0.011 0.011 0.011 

0.044 

0.047 

0.004 0.004 0.004 

0.005 

0.008 

0.014 0.014 0.013 

0.015 

0.015 

0.014 0.014 0.013 

0.015 

0.015 

Well 

EPA03 

EPA03-DUP 

Sampling Date

7/19/2005 

7/19/2005 

RPD 

0.063 39.0 13.8 3.35 

0.061 41.0 14.3 3.36 

3.2 5.0 3.6 0.3 

4.55 

5.71 

22.6 

12.4 

12.5 

0.8 

407 

408 

0.2 

0.80 

0.81 

1.4 

10.0 

10.2 

2.0 

38.9 

38.8 

0.3 

34.4 

34.8 

1.2 

0.97 

0.96 

1.0 

EPA01 

EPA01-DUP 

10/5/2005 

10/5/2005 

RPD 

0.286 52.1 20.8 5.73 

0.281 51.9 20.8 5.73 

1.8 0.4 0.0 0.0 

6.92 

6.90 

0.3 

14.7 

14.7 

0.0 

299 

299 

0.0 

2.96 

2.99 

1.0 

12.2 

12.2 

0.0 

28.4 

28.4 

0.0 

23.7 

22.3 

6.1 

4.19 

4.32 

3.1 

TR1 

TR1-DUP 

6/8/2006 

6/8/2006 

RPD 

0.028 73.1 21.9 5.31 

0.018 72.4 21.7 5.28 

43.5 1.0 0.9 0.6 

27.8 

25.2 

9.8 

15.5 

15.4 

0.6 

263 

261 

0.8 

0.021 

0.021 

0.0 

14.3 

14.4 

0.7 

18.1 

18.2 

0.6 

3.16 

3.57 

12.2 

10.5 

10.7 

1.9 

EPA06 

EPA06-DUP 

6/7/2006 

6/7/2006 

RPD 

0.147 42.8 16.7 4.57 

0.128 42.7 16.8 4.63 

13.8 0.2 0.6 1.3 

8.45 

8.43 

0.2 

14.0 

13.9 

0.7 

313 

318 

1.6 

1.19 

1.19 

0.0 

13.0 

13.0 

0.0 

34.0 

34.1 

0.3 

22.0 

22.2 

0.9 

8.77 

7.25 

19.0 

EPA08 EPA08-DUP 

9/18/2006 9/18/2006 

RPD 

0.249 51.5 20.3 5.12 0.235 50.1 19.8 5.13 

5.8 2.8 2.5 0.2 

6.87 6.77 

1.5 

13.1 13.0 

0.8 

220 227 

3.1 

2.28 2.25 

1.3 

11.5 11.5 

0.0 

27.3 27.5 

0.7 

19.3 18.1 

6.4 

5.60 5.17 

8.0 

EPA01 

EPA01-DUP 

9/25/2006 

9/25/2006 

RPD 

0.183 38.6 15.3 3.96 

0.181 40.0 15.8 4.11 

1.1 3.6 3.2 3.7 

4.76 

4.88 

2.5 

9.85 

9.91 

0.6 

153 

153 

0.0 

1.93 

1.99 

3.1 

9.03 

9.29 

2.8 

19.3 

20.2 

4.6 

15.5 

15.7 

1.3 

4.08 

4.40 

7.5 

TR0 

TR0-DUP 

7/19/2007 

7/19/2007 

RPD 

ND 3.68 7.29 0.015 

ND 3.73 7.56 0.020 

NA 1.3 3.6 28.6 

ND 

ND 

NA 

13.0 

13.0 

0.0 

202 

203 

0.5 

ND 

ND 

NA 

0.44 

0.43 

1.8 

0.134 

0.133 

0.7 

0.141 

0.117 

18.6 

ND 

ND 

NA 

T3C 

T3C-DUP 

7/26/2007 

7/26/2007 

ND 14.5 12.4 0.185 

ND 14.4 12.3 0.188 

ND 

ND 

12.4 

12.3 

245 

242 

ND 

ND 

2.12 

2.16 

0.724 

0.739 

1.53 

1.54 

ND 

ND 

RPD NA 0.7 0.8 1.6 NA 0.8 1.2 NA 1.9 2.1 0.7 NA 

54 

Table B1. Method reporting limits for selected analytes during five PRB sampling events and results of selected duplicate analyses.

RPD is the relative percent difference used to compare duplicate results, RPD = (|C2 - C1|)/((C1 + C2)/2) x 100, where C1 is the concentration in the s

concentration in the duplicate. The replicate results are in good agreement. Less than 5% difference is observed in 90 out 104 comparisons; better t

observed in 95 out of 104 observations. As expected, larger percent differences are observed in measurements approaching the detection limit and a

with the arsenic speciation analysis. 
Technique ICP-OES ICP-OES ICP-OES ICP-OES ICP-OES ICP-OES ICP-OES ICP-OES ICP-OES ICP-OES IC-ICP-MS IC-ICP-MS CE 

Date Al Ca Mg Mn Fe K Na Zn Si As As(III) As(V) Cl 

Field Blank 7/05 ND ND ND ND 0.02 0.06 ND ND ND ND ND ND ND 

Field Blank 10/05 ND ND ND ND ND ND 0.18 ND 0.55 0.007 ND ND 0.20 

Field Blank 6/06 ND 0.26 0.05 ND ND ND 0.70 ND 0.54 ND ND ND ND 

Field Blank 9/06 ND ND ND ND ND ND ND ND ND 0.005 ND ND ND 

Field Blank 7/07 ND ND ND ND ND ND ND ND ND ND ND ND ND 

Pump RinsePBTW-1 

8/02 ND 0.88 0.45 0.02 0.33 ND 1.85 ND 0.26 ND NA NA NA 

Pump RinseEPA05 

8/02 0.15 0.44 0.17 0.02 0.71 ND 0.57 ND 0.35 ND NA NA NA 

Table B2. Results of blank tests and pump rinsate tests.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 component

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 higmg/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 evaluate decontamination effectiveness. 

55 
Appendix C 

Table C1. Thermodynamic data for arsenic oxyanions. 

MINTEQA21 WATEQ4f2 LLNL3 

Nordstrom 

and Archer 

(2003) 

Raposo 

et al. (2004) 

ZakaznovaHerzog et al. 

(2006) 

Reaction 

As(III) species 

H3

AsO3(aq) = H2AsO3 

- + H+ -9.38 -9.36 -9.23 -9.15 -9.25 

H2AsO3 

- = HAsO3 2- + H+ -14.35 -14.70 -11.01 -14.70 

HAsO3 

2­ = AsO3 3- + H+ ND -15.70 ND -15.70 

As(V) species 

H3AsO4(aq) = H2AsO4 

- + H+ -2.24 -2.25 -2.25 -2.30 -2.25 

H2AsO4 

- = HAsO4 2- + H+ -7.01 -7.18 -6.75 -7.16 -7.06 

HAsO4 

2- = AsO4 3- + H+ -11.94 -11.79 -11.90 -11.65 -11.58 

Redox 

H3AsO4(aq) + H2(g) =

 H3AsO3(aq) + H2O(l) 22.96 23.40 22.56 19.35 

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 UCRLMA-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. Geochimica et Cosmochimica Acta, v. 70, p. 1928-1938. 

56 
Figure D1. Hydrograph of ground-water elevations 

measured at well DH-17. 

Figure D2. Hydrograph of ground-water elevations 

measured at well EPA02. 

Figure D3. Hydrograph of ground-water elevations 

measured at well EPA06. 

Figure D4. Hydrograph of ground-water elevations 

measured at well PBTW-2. 

Appendix D
 

Ground-Water Elevation Hydrographs 

57 
Figure D5. Hydrograph of ground-water elevations 

measured at well TR8. 

58 
Appendix E 

Potentiometric Surface in Vicinity of PRB 

Figure E1. Potentiometric surface at the water table in vicinity of PRB on June 19, 2002.
	

Figure E2. Potentiometric surface at the water table in vicinity of PRB on September 26, 2002.
	

59 
Figure E3. Potentiometric surface at the water table in vicinity of PRB on August 14, 2003.
	

60 

Figure E4. Potentiometric surface at the water table in vicinity of PRB on May 31, 2005.
	
Figure E5. Potentiometric surface at the water table in vicinity of PRB on July 20, 2005.
	

Figure E6. Potentiometric surface at the water table in vicinity of PRB on October 6, 2005.
	

61 
Figure E7. Potentiometric surface at the water table 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.
	
Figure E9. Potentiometric surface at the water table in vicinity of PRB on January 24, 2007.
	

Figure E10. Potentiometric surface at the water table in vicinity of PRB on July 18/19, 2007.
	

63 
Figure E11. Potentiometric surface at the water table in vicinity of PRB on October 1, 2007.
	

64 

Figure E12. Potentiometric surface at the water table in vicinity of PRB on April 1, 2008.
	
Figure E13. Potentiometric surface at the water table in vicinity of PRB on June 24, 2008.
	

65 

Office of Research and Development (8101R) 

Washington, DC 20460 

Official Business 

Penalty for Private Use 

$300 

PRESORTED STANDARD
 

POSTAGE & FEES PAID
 

EPA
 

PERMIT NO. G-35