4>EPA
United States
Environmental Protection
Agency
Capstone Report on the
Application, Monitoring, and
Performance of Permeable
Reactive Barriers for
Ground-Water Remediation:
Volume 1
Performance Evaluations at Two Sites
Ground Water Flow
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EPA/600/R-03/045a
August 2003
on the
of
for
Volume 1 - Performance Evaluations at Two
Richard T. Wilkin and Robert W. Puls
Ground Water and Ecosystems Restoration Division
National Risk Management Research Laboratory
Ada, Oklahoma 74820
National Risk Management Research Laboratory
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, OH 45268
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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.
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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 technologi-
cal and management approaches for preventing and reducing risks from pollution that threatens human health and
the environment. The focus of the Laboratory's research program is on methods and their cost-effectiveness for
prevention and control of pollution to air, land, water, and subsurface resources; protection of water quality in public
water systems; remediation of contaminated sites, sediments and ground water; prevention and control of indoor air
pollution; and restoration of ecosystems. NRMRL collaborates with both public and private sector partners to foster
technologies that reduce the cost of compliance and to anticipate emerging problems. NRMRL's research provides
solutions to environmental problems by: developing and promoting technologies that protect and improve the
environment; advancing scientific and engineering information to support regulatory and policy decisions; and
providing the technical support and information transfer to ensure implementation of environmental regulations and
strategies at the national, state, and community levels.
This publication has been produced as part of the Laboratory's strategic long-term research plan. It is published and
made available by EPA's Office of Research and Development (ORD) to assist the user community and to link
researchers with their clients. The purpose of this document is to provide detailed performance monitoring data on
full-scale Permeable Reactive Barriers (PRBs) installed to treat contaminated ground water at two different sites.
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 by chlorinated solvents, chromium, arsenic, nitrates, and other
organic and inorganic compounds and are considering the use of this cost-effective technology. The PRBs
discussed in this report are among the oldest full-scale systems available for study and provide an opportunity to
analyze the performance of systems with more than five years of field history. In addition, the PRBs examined here
have contrasting design and hydrogeochemical characteristics that are useful in the context of gaining insight about
the factors that govern PRB longevity and long-term performance. 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.
Ground Water and EcosystemsKestoration Division
National Risk Management Re/earcn Laboratory
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Research results discussed in this report explore the geochemica! and microbiological processes within zero-valent
iron Permeable Reactive Barriers (PRBs) that may contribute to changes through time in iron reactivity and
decreases in reaction zone permeability. Two full-scale PRBs were evaluated in this study: the U.S. Coast Guard
Support Center PRB located near Elizabeth City, North Carolina, and the Denver Federal Center PRB in Lakewood,
Colorado. Detailed water sampling and analysis, core sampling, and solid-phase characterization studies were
carried out to: i) evaluate spatial and temporal trends in contaminant concentrations and key geochemical
parameters; ii) characterize the type and nature of surface precipitates forming overtime in the reactive barriers; and
iii), identify the type and extent of microbiological activity within and around the reactive barriers.
Trends in geochemical parameters (e.g., pH and oxidation-reduction potential) may signal changes in system
performance, but no clear correlations between these parameters and decreased system performance have been
observed to date at the sites studied. Long-term trends in geochemical parameters are consistent with contaminant
removal trends observed at both sites. Spatial and temporal variations in the concentration distribution of terminal
electron accepting species (e.g., sulfate), specific conductance, and Eh suggest that both anaerobic iron corrosion
and microbial activity play important roles in controlling the oxidation-reduction potential in iron barriers. Low Eh
values (<100 mV relative to the standard hydrogen electrode) and decreases in the specific conductance of ground
water between upgradient contaminant plumes and sampling points within reactive iron media are consistently
observed in normally operating PRB systems.
The rate of mineral and biomass buildup was evaluated at both sites. The principal factors that determine the amount
of mineral precipitation in zero-valent iron PRBs are flow rate, ground-water chemistry, and microbial activity. After
five years of operation, the Elizabeth City and Denver Federal Center reactive barriers have developed consistent
patterns of spatially variable mineral precipitation and microbial activity. The development of precipitation and
biomass fronts result from abrupt geochemical changes that occur at upgradient interface regions coupled with
ground water solute transport. Upgradient regions at both sites investigated in this study have witnessed the
greatest accumulation of mineral mass and biomass. However, neither of the sites of this study show complete filling
of available pore space after five years, suggesting that flow characteristics should not be affected by the
accumulation of authigenic components. For zero-valent iron systems, the reactive media is a long-term sink for C,
S, Ca, Si, Mg, and N. Porosity loss in the iron media due to precipitation of inorganic carbon and sulfur minerals can
be estimated by integrating the concentrations of inorganic carbon and sulfur as a function of distance in the iron and
estimating the volume loss by using the molar volumes of zero-valent iron, calcium carbonate, iron carbonate, and
iron sulfide. Porosity loss estimates have ranged from about 1% to 4% per year in this study. Based on these
estimates, the average porosity of the PRB at Elizabeth City, for example, would not be expected to approach that
of the surrounding aquifer for 15 to 30 years. As corrosion minerals form on the surface of the iron media, reactive
surfaces are coated, presumably decreasing the effective reactive surface area. However, corrosion products
formed include some minerals which themselves are highly reactive and capable of transforming inorganic and
organic contaminants into immobile or non-toxic forms. This phenomenon must also be factored into lifetime
projections.
While long-term performance observations of the Elizabeth City and Denver Federal Center site are now past five
years, there has still not been sufficient time to adequately predict the lifetime of these PRBs or most other PRBs. It
is clear that lifetimes exceeding 10 years are reasonable to expect under some conditions and that PRBs may
function adequately for much longer. Continued studies are needed to better predict longevity based on ground-
water composition, flow rate and contaminant flux.
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of
Notice ii
Foreword Ill
Abstract iv
Table of Contents v
Figures vii
Tables xii
Acknowledgments xiv
1.0 Introduction 1
1.1 Federal Tri-Agency Initiative 2
1.1.1 DoD Studies 2
1.1.2 DOE Studies 3
1.1.3 EPA Studies 3
2.0 Site Descriptions 5
2.1 U.S. Coast Guard Support Center 5
2.2 Denver Federal Center 5
3.0 Elizabeth City PRB Monitoring Results 9
3.1 Ground Water Monitoring 9
3.1.1 Contaminant Behavior 9
3.1.2 Geochemical Parameters 13
3.1.3 Dissolved Cations and Anions 18
3.2 Core Sampling at Elizabeth City 25
3.2.1 Carbon Analysis 25
3.2.2 Sulfur Analysis 33
3.2.3 Cr Extractions 39
3.2.4 X-ray Diffraction Analysis 39
3.2.5 Scanning Electron Microscopy 41
3.2.6 Microbial Characterization 46
3.3 Summary of Results from the Elizabeth City Site 52
4.0 Denver Federal Center Monitoring Results 53
4.1 Ground Water Monitoring 53
4.1.1 Gate 1 Contaminant Behavior 53
4.1.2 Gate 2 Contaminant Behavior 58
4.1.3 Gate 3 Contaminant Behavior 58
4.1.4 Geochemical Parameters 58
4.1.5 Hydrogen Gas Concentrations 62
4.1.6 Dissolved Cations and Anions 64
4.2 Core Sampling at the Denver Federal Center 68
4.2.1 Carbon Analysis 71
4.2.2 Sulfur Analysis 71
4.2.3 X-ray Diffraction Analysis 71
4.2.4 Scanning Electron Microscopy 71
4.2.5 Microbial Characterization 80
4.3 Summary of Results from the Denver Federal Center Site 82
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5,0 Factors Affecting Longevity and Performance 87
5.1 Fe° Dissolution 87
5.2 Anion Composition 89
5.2.1 Bicarbonate Reactions 89
5.2.2 Sulfate Reactions 96
5.2.3 Nitrate Reactions 97
5.2.4 Reactions with Silica 97
5.2.5 Reactions with Oxygen 98
5.3 Mineral Precipitation 99
5.3.1 Pore Volume Reduction 99
5.3.2 Loss of Reactivity 103
5.4 Microbial Activity 104
5.5 Hydrogeologica! Issues 107
6.0 State of Permeable Reactive Barrier Technology and
Lessons Learned from Long-Term Performance Monitoring 111
6.1 Permeable Reactive Barriers: An Accepted Remedial
Option for Containment & Treatment of Contaminated Ground Water 111
6.2 Lessons Learned: Site Characterization and PRB Construction 112
6.3 Lessons Learned: Long-term Performance Assessments of PRBs 112
6.3.1 Recommendations for Future Research 113
7.0 References 115
Appendix A 121
Appendix B 135
VI
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Figure 2.1 Plan view map of the PRB at the U.S. Coast Guard Support
Center, Elizabeth City, NC 7
Figure 2.2 Plan view map of the PRB at the Denver Federal Center,
Lakewood, CO (after McMahon et al., 1999) 8
Figure 2.3 Schematic cross-section of the Denver Federal Center
funnel-and-gate system (after McMahon et al., 1999) 8
Figure 3.1 Concentrations of contaminants through time in monitoring
wells located hydraulically upgradient of the Elizabeth City PRB 10
Figure 3.2 Concentrations of TCE (ng/L) through time in monitoring
wells located hydraulically downgradient of the Elizabeth City PRB 13
Figure 3.3 Cross-sectional profiles showing total chromium
concentrations (mg/L) in transects 1, 2, and 3 (Elizabeth City PRB) 14
Figure 3.4 Cross-sectional profiles showing TCE concentrations
(ng/L) in transects 1, 2, and 3 (Elizabeth City PRB) 15
Figure 3.5 Cross-sectional profiles showing cis-DCE concentrations
(ng/L) in transects 1, 2, and 3 (Elizabeth City PRB) 16
Figure 3.6 Cross-sectional profiles showing VC concentrations
(u,g/L) in transects 1, 2, and 3 (Elizabeth City PRB) 17
Figure 3.7 Cross-sectional profiles showing pH distributions
in transects 1, 2, and 3 (Elizabeth City PRB) 19
Figure 3.8 Cross-sectional profiles showing Eh distributions
(mV) in transects 1, 2, and 3 (Elizabeth City PRB) 20
Figure 3.9 Cross-sectional profiles showing specific conductance
distributions (jiS/cm) in transects 1, 2, and 3 (Elizabeth City PRB) 21
Figure 3.10 Cross-sectional profiles showing calcium concentrations
(mg/L) in transects 1, 2, and 3 (Elizabeth City PRB) 22
Figure 3.11 Cross-sectional profiles showing magnesium concentrations
(mg/L) in transects 1, 2, and 3 (Elizabeth City PRB) 23
Figure 3.12 Cross-sectional profiles showing sodium concentrations
(mg/L) in transects 1, 2, and 3 (Elizabeth City PRB) 24
Figure 3.13 Cross-sectional profiles showing potassium concentrations
(mg/L) in transects 1, 2, and 3 (Elizabeth City PRB) 26
Figure 3.14 Cross-sectional profiles showing chloride concentrations
(mg/L) in transects 1, 2, and 3 (Elizabeth City PRB) 28
Figure 3.15 Cross-sectional profiles showing sulfate concentrations
(mg/L) in transects 1, 2, and 3 (Elizabeth City PRB) 29
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Figure 3.16 Cross-sectional profiles showing alkalinity distributions
(mg/L) in transects 1, 2, and 3 (Elizabeth City PRB) 30
Figure 3.17 Cross-sectional profiles showing nitrate concentrations
(mg/L) in transects 1, 2, and 3 (Elizabeth City PRB) 31
Figure 3.18 Cross-sectional profiles showing silica concentrations
(mg/L) in transect 2 (Elizabeth City PRB) 33
Figure 3.19 Coring locations and monitoring well locations at
the Elizabeth City PRB (plan view) 34
Figure 3.20 Cross-sectional profile showing concentration distribution
of inorganic carbon in the solid phase (|j.g/g=ppm), Elizabeth City
PRB (June 2002) 36
Figure 3.21 Concentrations of inorganic carbon (u.g/g) in core materials
through time, Elizabeth City PRB 37
Figure 3.22 Cross-sectional profile showing concentration distribution
of sulfur in the solid phase (|ig/g=ppm), Elizabeth City PRB (June 2002) 38
Figure 3.23 Powder X-ray diffraction data from fine-grained materials
removed via sonication from cores collected at the
Elizabeth City PRB: a) core EC060300-4; b) core EC050801-3 40
Figure 3.24 Scanning electron micrographs of samples from the
Elizabeth City PRB: a) sample EC060300-4-1; b) sample
EC060300-4-3; c) EC060300-4-7 42-44
Figure 3.25 Iron concentration versus oxygen concentration in iron
grains and surface precipitates (SEM-EDX) 46
Figure 3.26 Element concentrations in surface precipitates from the
Elizabeth City PRB 48
Figure 3.27 Cross-sectional profile showing concentration distribution
of biomass (from PLFA data) in picomoles per gram,
Elizabeth City PRB (June 2002) 49
Figure 3.28 Histograms of microbial biomass concentrations
(from PLFA data) in picomoles per gram in aquifer materials,
iron from near the upgradient aquifer/iron interface, and iron
from near the downgradient aquifer/iron interface 50
Figure 3.29 Pie graphs showing structural distribution of PLFA
compounds (average values) at the Elizabeth City site 51
Figure 4.1 Coring locations and monitoring well locations at the
Denver Federal Center, gate 1 (plan view) 54
Figure 4.2 Coring locations and monitoring well locations at the
Denver Federal Center, gate 2 (plan view) 55
Figure 4.3 Coring locations and monitoring well locations at the
Denver Federal Center, gate 3 (plan view) 56
Figure 4.4 Concentrations of contaminants through time in monitoring
wells from the Denver Federal Center, gate 1 (data from FHWA):
a) well GSA-21 (upgradient); b) well C1-I1 (iron wall);
c) well GSA-20 (downgradient) 57
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Figure 4.5 Concentrations of contaminants through time in monitoring
wells from the Denver Federal Center, gate 2 (data from FHWA):
a) well GSA-26 (upgradient); b) well C2-I2 (iron wall);
c) well GSA-25 (downgradient) 59
Figure 4.6 Depth-resolved concentrations of a) contaminants (|o,g/L)
and b) sulfate, calcium, and iron (mg/L) in wells GSA-26
and GSA-25 from the Denver Federal Center (gate 2) 60
Figure 4.7 Concentrations of contaminants through time in monitoring
wells from the Denver Federal Center, gate 3 (data from FHWA):
a) well GSA-31 (upgradient); b) well C3-12 (iron wall);
c) well GSA-30 (downgradient) 61
Figure 4.8 Average pH values through time in wells from upgradient,
iron wall, and downgradient positions relative to gate 1,
gate 2, and 3 at the Denver Federal Center 62
Figure 4.9 Average specific conductance values (u.S/cm) through time
in wells from upgradient, iron wall, and downgradient
positions relative to gate 1, gate 2, and gate 3 at the
Denver Federal Center 63
Figure 4.10 Average Eh (V) values through time in wells from upgradient,
iron wall, and downgradient positions relative to gate 1,
gate 2, and gate 3 at the Denver Federal Center 63
Figure 4.11 Concentrations of dissolved hydrogen (log molar) as
a function of sampling position and time in gate 1 at
the Denver Federal Center. Also shown are the
concentration ranges of dissolved hydrogen measured in
the iron media in gate 2 and gate 3 64
Figure 4.12 Average (± 1 s.d.) concentrations of Ha, K, Ca, Mg, sulfate,
bicarbonate, chloride, and silica (mg/L) as a function of
sampling position in gate 1 at the Denver Federal Center 65
Figure 4.13 Average (± 1 s.d.) concentrations of Na, K, Ca, Mg, sulfate,
bicarbonate, chloride, and silica (mg/L) as a function of
sampling position in 2 at the Denver Federal Center 66
Figure 4.14 Average (± 1 s.d.) concentrations of Na, K, Ca, Mg, sulfate,
bicarbonate, chloride, and silica (mg/L) as a function of
sampling position in gate 3 at the Denver Federal Center 67
Figure 4.15 Picture showing cemented nodules recovered from core
collected at the Denver Federal Center 69
Figure 4.16 Picture showing the appearance of cores collected at
the Denver Federal Center, from gate 2 near the upgradient
peagravel/iron interface 69
Figure 4.17 Concentration distribution of solid phase inorganic
carbon in angle cores collected from 1 at the Denver
Federal Center 72
Figure 4.18 Concentration distribution of solid phase inorganic
carbon in angle cores collected from gate 2 at the Denver
Federal Center 72
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Figure 4.19 Concentration distribution of solid phase inorganic
carbon in a vertical core collected from gate 2 at the
Denver Federal Center 73
Figure 4.20 Concentration distribution of solid phase sulfur
in angle cores collected from gate 1 at the Denver
Federal Center 73
Figure 4.21 Concentration distribution of solid phase sulfur
in angle cores collected from gate 2 at the Denver
Federal Center 74
Figure 4.22 Inorganic carbon concentrations versus total sulfur
a) Elizabeth City core materials; b) Denver Federal
Center 75
Figure 4.23 Powder X-ray diffraction data from fine-grained materials
removed via sonication from cores collected at the
Denver Federal Center PRB (core C2-3-71801) 76
Figure 4.24 Scanning electron micrographs of samples from the
Denver Federal Center PRB:a) sample C2-17-71300-2;
b) sample C2-17-71300-7; c) sample C1-2-71000-3 77-79
Figure 4.25 Concentration of microbial biomass (from PLFA data)
in picomoles per gram in iron from near the upgradient
peagravel/iron interface and iron from near the downgradient
peagravel/iron interface: a) gate 1; b) gate 2 82
Figure 4.26 Concentration distribution of solid phase sulfur and microbial
biomass (from PLFA data) in a vertical core collected from gate 2
at the Denver Federal Center (vertical cores C2-1-71901,
C2-2-71901, and C2-3-71901) 83
Figure 4.27 Pie graphs showing average structural distribution of PLFA
compounds in core materials from the Denver Federal Center 84
Figure 5.1 Redox -pH diagram showing composition of ground water
from the Elizabeth City iron wall compared to equilibrium
trends for the Fe°-Fe(OH)3> Fe°-Fe3O4, and Fe2+-Fe(OH)3 couples 88
Figure 5.2 Redox-pH diagram for the Fe-H2O system at 25 eC, showing
speciation of iron (dashed lines) and stability fields of
iron-bearing minerals (solid lines) 90
Figure 5.3 Redox-pH diagram for the Fe-CO2-H2O system at 25 BC,
showing speciation of iron (dashed lines) and stability fields of
iron-bearing minerals (solid lines) 90
Figure 5.4 Redox-pH diagram for the Fe-S-CO2-H2O system at 25 SC,
showing speciation of iron (dashed lines) and stability fields of
iron-bearing minerals (solid lines) 91
Figure 5.5 Redox-pH diagram for the Fe-S-CO2-H2O system at 25 aC,
showing speciation of iron (dashed lines) and stability fields of
iron-bearing minerals (solid lines) 91
Figure 5.6 Redox-pH diagram for the Fe-CO2-H2O system at 25 eC,
showing speciation of iron (dashed lines) and stability fields of
iron-bearing minerals (solid lines) 92
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Figure 5.7 Redox-pH diagram for the Fe-CO2-H2O system at 25 aC,
showing speciation of iron (dashed lines) and stability fields of
iron-bearing minerals (solid lines) 92
Figure 5.8 Upgradient ground-water compositions (molar ratios)
and TDS values for the Elizabeth City and Denver
Federal Center PRB sites 93
Figure 5.9 Comparison of total dissolved solids concentrations at
PRB sites studied in the Tri-Agency initiative 93
Figure 5.10 Solubility diagram showing the stability field of carbonates
as a function of pH and log activities of Ca, Fe, Mg, and
dissolved inorganic carbon compared to ground-water
compositions from upgradient, iron wall, and
downgradient sampling locations (Elizabeth City PRB) 95
Figure 5.11 Saturation indices of magnesium-bearing phases (brucite,
Mg(OH)2; sepiolite, Mg4(OH)2Si6O15-H2O) as a function of pH
in ground water from upgradient, iron wall, and downgradient
sampling locations (Elizabeth City PRB) 98
Figure 5.12 Conceptual model of the impact of mineral and
biomass accumulation to PRB hydraulic performance 99
Figure 5.13 Fractional porosity reduction as a function of inorganic
carbon concentration in the solid phase 100
Figure 5.14 Fractional porosity reduction as a function of sulfur
concentration in the solid phase 102
Figure 5.15 Fractional porosity reduction as a function of the positive
molar volume change as iron metal reacts to form
magnetite, hematite, goethite, and ferrihydrite 102
Figure 5.16 Concentration versus time in batch tests: a) chromium; b) TCE 105
Figure 5.17 Comparison of PLFA distribution in four iron walls 107
Figure 5.18 Water levels in Elizabeth City monitoring wells 109
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Table 2.1 Comparison of PRBs investigated in this study 6
Table 3.1 Contaminant concentrations in ground water upgradient
of the Elizabeth City PRB 11
Table 3.2 Contaminant concentrations in ground water downgradient
of the Elizabeth City PRB 12
Table 3.3 Geochemical parameters in upgradient, iron wall, and
downgradient locations, Elizabeth City PRB 18
Table 3.4 Changes in concentrations of metals in transect 2 as a
function of time and depth below ground surface 27
Table 3.5 Changes in concentrations of anions in transect 2 as a
function of time and depth below ground surface 32
Table 3.6 Cores collected for analysis at the Elizabeth City PRB 35
Table 3.7 Results of powder X-ray diffraction analysis of core
materials from the Elizabeth City PRB 41
Table 3.8 Pearson's correlation matrix of element concentrations
determined by SEM-EDX analysis 46
Table 3.9 Summary of PLFA data from the Elizabeth City PRB 47
Table 4.1 Cores collected for analysis at the Denver Federal
Center PRB 70
Table 4.2 Results of powder X-ray diffraction analysis of core
materials from the Denver Federal Center PRB 76
Table 4.3 Results of SEM-EDX analysis of core materials from
The Denver Federal Center PRB 80
Table 4.4 Summary of PLFA data from the Denver Federal Center PRB 81
Table 5.1 Mineral precipitates identified in iron walls 94
Table 5.2 Molar volume and density of mineral precipitates 101
Table 5.3 Core samples used in batch reactivity tests 103
Table 5.4 Water compositions used in batch reactivity tests 104
Table 5.5 Summary of rate data for reactions of TCE and 1,1,1-TCA
with zero-valent iron (unreacted and from field PRBs) 106
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Appendix A-1: Inorganic carbon and sulfur concentrations in cores
collected from the Elizabeth City and Denver Federal
Center PRB sites 122-126
Appendix A-2: Reduced Sulfur Speciation in Elizabeth City and
Denver Federal Center Cores 127
Appendix A-3: Inorganic Carbon and Sulfur Concentrations in
Denver Federal Center Cores 128-133
Appendix B-1: Phospholipid fatty-acid (PLFA) extract data from cores
collected from the Elizabeth City and Denver Federal
Center PRB sites 136-140
XIII
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The authors would like to acknowledge all of the participants of the Tri-Agency PRB research initiative, in particular
Arun Gavaskar (Battelle), Bruce Sass (Battelle), Neeraj Gupta (Battelle), Woong-Sang Yoon (Battelle), Libby West
(Oak Ridge National Laboratory), Nic Korte (Oak Ridge National Laboratory), Liyuan Liang (Cardiff University), and
Matt Turner (Interstate Technologies Regulatory Cooperation). Members of the Permeable Reactive Barriers
Research Team contributed to the research described in this report, especially S. Acree, F. Beck, P. Clark, K. Jones,
M. McNeil, C. Paul, and C. Su. Mantech Environmental Research Services Corporation provided analytical support
both in the field and in the laboratory, T. Sivavec (General Electric Corporate) provided X-ray photoelectron
spectroscopic analysis of iron core materials. We thank R. Ford (U.S. EPA) for many discussions throughout this
study and D. Myers (East Central University) for the scanning electron microscopy. Shim Myung-Hwa, a visiting
student from the Kwangju Institute of Science and Technology (South Korea), is acknowledged for her help on parts
of this study. Martha Williams (CSC) provided support in document preparation. J.P. Messier (U.S. Coast Guard
Support Center) is thanked for providing site assistance at the Elizabeth City site and C. Eriksson (Pacific Western
Technologies, Ltd.), J. Jordon (Pacific Western Technologies, Ltd.), and M. Gasser (Pacific Western Technologies,
Ltd.) are thanked for site assistance at the Denver Federal Center site. We also acknowledge the U.S. Coast Guard
for access to the Elizabeth City site and the Federal Highway Administration and General Services Administration for
access to the Denver Federal Center site. Reviews of the document were provided by Eric Reardon (University of
Waterloo), Liyuan Liang (Cardiff University), Steve Shoemaker (Dupont), Robert Powell (Powell and Associates),
Carl Eriksson (Pacific Western Technologies, Ltd.), and Thomas Holdsworth (U.S. EPA); their thoughtful comments
are appreciated.
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1.0
The permeable reactive barrier (PRB) technology has gained acceptance as an effective ground-water remediation
strategy for the treatment of a variety of chlorinated organic and inorganic compounds. The technology combines
subsurface fluid-flow management with contaminant treatment by chemical, physical or biological processes, or by
combinations of these three principal process categories. The PRB methodology has advantages over traditional pump-
and-treat systems in that it is passive and a large plume can be treated in a cost-effective manner. More than one
hundred implementations of the technology worldwide have proven that passive reactive barriers can be cost-effective
and efficient approaches to remediate a variety of compounds of environmental concern. Yet, few case studies are
available that evaluate the long-term performance of these in-situ systems, especially with respect to the long-term
efficiency of contaminant removal, the buildup of mineral precipitates, and the buildup of microbial biomass (e.g.,
O'HannesinandGillham, 1998; McMahon etal., 1999; Pulsetal., 1999a; Vogan, 1999; Phillips etal., 2000; Gavaskaret
al., 2002; Liang et al., 2002; Wilkin et al., 2002).
Granular iron is most often used as a reactive media in full-scale site remediation efforts using the PRB technology. The
prevalent use of zero-valent iron mainly stems from its low cost (approx. $350 to $450 per ton in 2002), availability, and
documented ability to degrade a wide variety of contaminant types. In the of chlorinated volatile organic
compounds such as perchloroethylene (PCE) and trichloroethylene (TCE), contaminant removal by zero-valent iron may
be due to direct electron transfer, reaction with Fe2+ produced during anaerobic iron corrosion, or due to catalytic
hydrogenation reactions (e.g., Matheson and Tratnyek, 1994; Gillham and O'Hannesin, 1994; Johnson and Tratnyek,
1994; Roberts etal., 1996; Orth and Gillham, 1996; Fennelly and Roberts, 1998). In the case of inorganic contaminants
such as U, Cr, and As, contaminant removal may be achieved through reductive precipitation or adsorption (e.g., Cantrell
etal., 1995; Blowesetal., 1997,2000; Powell etal., 1998; Fiedor etal., 1998; Lackovicet al., 2000; Morrison etal., 2001,
2002). New insight regarding the reactive behavior of zero-valent iron and a more detailed understanding of reaction
kinetics and reaction pathways involving zero-valent metals continues to emerge from laboratory studies (e.g., Burris et
al., 1995; Agrawal and Tratnyek, 1996; WQst et al., 1999; Deng et al., 1999; Su and Puls, 1999; Nam and Tratnyek, 2000;
Schlicker et al., 2000; Arnold and Roberts, 2000; Farrell et al., 2000; Ruiz et al., 2000; Su and Puls, 2001; Melitas et al.,
2001; Scherer et al., 1998; Lien and Wilkin, 2002; Alowitz and Scherer, 2002; Kober et al., 2002).
In addition to reaction processes involving contaminant species, zero-valent iron also impacts the biogeochemical
behavior of the typically more concentrated assortment of ground-water solutes. Reaction processes that involve the
major anionic (e.g., Cl~, SO42~, HCO3~) and major cationic ground-water components (e.g., Ca2+, Mg2+) govern the kinetics
and pathways of iron corrosion, mineral precipitation, microbial activity, and gas production within and around the
reactive media (e.g., Reardon, 1995). The cumulative effect of these processes through time can lead to changes in the
reactivity, porosity, and hydraulic permeability of a PRB.
Research results described in this report explore the geochemical and microbiological processes occurring within zero-
valent iron PRBs that may contribute to changes in iron reactivity and decreases in reaction zone permeability that, in
turn, may eventually lead to system failure or plugging. Field studies were carried out at two geographically,
hydrogeologically, and geochemically distinct iron barrier installation sites (U.S. Coast Guard Support Center, Elizabeth
City, NC; and Denver Federal Center, Lakewood, CO). The research approach consisted of intensive ground-water
sampling, mineralogical and microbiological characterization of core materials, and geochemical modeling to compare
expected trends in the type and mass of mineral precipitates with observations from the field. Specific objectives of this
research project were to:
1) Evaluate spatial and temporal trends in contaminant concentrations and key geochemical parameters that
may impact reactivity and steady system performance.
2) Characterize the type and properties of surface precipitates forming over time at the upgradient aquifer/iron
interface, within the iron media, and at the downgradient/iron interface.
3) Develop conceptual models that predict the type and rate of precipitate formation based on iron characteristics
and water chemistry.
-------
4) Identify the type and extent of microbiological activity at the upgradient aquifer/iron interface, within the iron
media, and at the downgradient/iron interface.
5) Define practical and cost-effective protocols for long-term performance assessments at permeable reactive
barrier installations.
A detailed analysis of the rate of surface precipitate buildup in PRBs is critical for understanding how long these systems
will remain effective and what methods may be employed to extend their lifetime or improve their performance (e.g.,
Geiger et al., 2002). Different types of minerals and surface coatings have been observed to form under different
geochemical conditions that would appear to be dictated by aquifer chemistry and the composition of the permeable
reaction zone (Powell etal., 1995; Mackenzie etal., 1999; Pulsetal., 1999b; Liang etal., 2000; Phillips etal., 2000; Bonin
et al., 2000; Roh et al., 2000; Wilkin et al., 2002; Furukawa et al., 2002). Furthermore, microbiological impacts are also
important to understand in order to better predict how long these systems will remain effective in the subsurface
(Matheson, 1994; Weathers et al., 1997; Till et al., 1998; Gu et al., 1999, 2002; Scherer et al., 2000; Gandhi et al., 2002).
The presence of a large reservoir of iron coupled with plentiful substrate availability in the form of hydrogen supports the
metabolic activity of iron-reducing, sulfate-reducing, and/or methanogenic bacteria. This enhanced microbial activity
may beneficially influence zero-valent iron reductive dehalogenation reactions through favorable impacts to the iron
surface or through direct microbial transformations of the target compounds. However, this enhancement may come at
the expense of faster corrosion leading to faster precipitate buildup and potential biofouling of the permeable treatment
zone.
1.1
Research described in this report was carried out as part of a Tri-Agency cooperative effort between the United States
(U.S.) Department of Defense (DoD), the U.S. Department of Energy (DOE), and the U.S. Environmental Protection
Agency (EPA). This collaborative initiative allowed the three federal agencies to leverage technical and funding
resources and to share experiences at over 10 PRB installations across the U.S. The primary goal of the research
initiative was to evaluate the longevity and hydraulic performance of zero-valent iron PRBs in various hydrogeological
and geochemical settings. Members of the Tri-Agency initiative met periodically and conducted regular conference calls
to discuss research progress. Results of research projects conducted by members of the Tri-Agency effort are reported
in Gavaskar et al. (2002), Liang et al. (2002), and Wilkin et al. (2002). A combined final product outlining conclusions of
the Tri-Agency study is expected in 2003.
This report provides a detailed exploration of long-term monitoring results obtained over the initial 5-year operation
period of PRBs at the U.S. Coast Guard Support Center (Elizabeth City, NC) and the Denver Federal Center (Lakewood,
CO). Results of the study are presented in two volumes. Volume 1 (this volume) presents a performance assessment
and data on the geochemical and microbiological factors that impact the performance of zero-valent iron PRBs. Volume
2 presents ground water monitoring practices and procedures employed in this study as well as soil core collection,
preservation, and analysis methods. The remainder of this section presents the general scope and major conclusions of
the DoD, DOE, and EPA contributions to the Tri-Agency collaborative effort.
1.1.1
Results of the DoD study on PRB longevity and hydraulic performance are reported in Gavaskar et al. (2002). Field data
were collected and analyzed from PRBs at several DoD sites. The longevity evaluation focused primarily on PRBs at the
Moffett Field former Naval Air Station (CA) and the former Lowry Air Force (CO). Both PRBs have a funnel-and-
gate design and were installed prior to 1996. The longevity evaluation consisted of geochemical monitoring and
modeling of ground water, iron core collection and analysis, and accelerated column tests. Hydraulic performance was
also studied at the Seneca Army Depot (NY, installed 1998) and the Dover Air Force Base (DE, installed 1997).
Methodologies used for the hydraulic performance testing were water level measurements and a variety of available in-
situ flow sensors.
Five years after installation of the Moffett Field PRB, concentrations of TCE, PCE, and cis-1,2 dichloroethylene (cis-
DCE) in effluent ground water from the reactive cell were all below their respective maximum concentration levels
(MCLs). Treatment of contaminants occurred mainly in the upgradient region of the reactive media. As of 2001, a clean
front of ground water had not been identified in the downgradient aquifer, although there was an indication that a clean
front would occur in the future. Similar results with respect to contaminant treatment were observed at the former Lowry
Air Force (AFB). At the Moffett Field and Lowry AFB sites, concentrations of dissolved calcium, iron, magnesium,
sulfate, nitrate, alkalinity, silica, and total dissolved solids flowing to the PRBs were significantly reduced in effluent
ground water compared to influent ground water. At both sites, pH values within the reactive media rose to levels as high
as 11.5 and ORP values dropped to as low as -821 mV. Solid-phase characterization studies were carried out to
evaluate the mineralogy of precipitates that formed in the reactive media.
-------
Careful and periodic water level measurements were found to give the best results with respect to hydraulic performance
monitoring. In-situ, direct-flow measurements in some cases gave flow direction results that were contradictory from
those indicated by water level measurements. This was thought to be due to the fact that in-situ techniques were point
estimates and more indicative of localized flow conditions, whereas water level trends are more indicative of the overall
flow regime around the PRBs studied. Estimated effective residence times varied from about 9 days at the Moffett Field
site to about 25 days at the Lowry AFB.
1.1.2
Results of the DOE study are presented in two reports {Liang et al., 2002; Moline et al., 2002) and recent publications
(e.g., Kamolpornwijit et al., 2003; Liang et al., 2003). The DoE study focused on PRBs in Monticello, Utah and at the Y-
12 Pathway-2 in Oak Ridge, Tennessee. The Monticello site was a former vanadium and uranium ore-processing mill.
A zero-valent iron funnel-and-gate system was installed in July 1999 to treat uranium-contaminated ground water. The
Y-12 Pathway-2 PRB was installed in November 1997 as a trench that was intended to capture shallow ground-water
flow contaminated with uranium. The DOE study consisted of geochemical and hydrogeologic monitoring, core
sampling, column tests, and geochemical modeling.
Slug tests at the Monticello barrier in wells immediately upgradient of the PRB resulted in hydraulic conductivity values
of 3.1 to 27 m/d. Colloidal borescope measurements showed particle velocities ranging from 4.3 to 43 m/day. The
borescope measurements showed evidence of vertical variability in flow conditions within each of the wells tested. A
multiparameter tracer test conducted one year after the PRB was installed resulted in upgradient seepage velocities of
0.36 to 3.6 m/d. Tracer tests were conducted by injecting anionic tracers (bromide, iodide) or dissolved gas tracers
(helium, neon, argon), and monitoring their appearance in a network of adjacent monitoring wells. Within the PRB at
Monticello, considerable lateral transport was observed.
The DOE report indicates that flow directions and velocities are dependent on the type of tracer test used. A general
conclusion presented was that an analysis of water levels provides information about average gradients but may be
difficult to interpret within the context of site heterogeneity. Potentiometric surfaces have the most value for regional
conceptualization of water flow patterns and for delineating gross features such as ground-water mounding. On the
other hand, tracer tests provide definitive results on a more local scale, but the scale of the measurement needs to be
considered when interpreting such data or when extrapolating trends to adjacent aquifer regions (Moline et al., 2002).
Liang et al. (2002; 2003) used the geochemical equilibrium model PHREEQC to evaluate mineral saturation indices in
waters from several sites including the Monticello and Y-12 Plant PRBs. This analysis was carried out to understand the
types and quantities of minerals that might form as water passes through and chemically equilibrates with the reactive
medium. The results of the modeling show that the buffering capacity and flow rate of the influent ground water is
important in determining the equilibrium pH in the Fe° media and in effluent water from Fe° columns. The predicted
spatial distribution of secondary minerals based on pore water chemistry provides a direct indication of changing flow
characteristics over time and has been shown to be in agreement with the results of tracer testing at the Y-12 site.
Column tests were carried out at the Y-12 Plant site to examine iron deterioration processes, changes to influent water
chemistry after reaction with iron metal, and mineral precipitation processes. Column tests were run for over 14 months.
In the column experiments, in-situ ground water was used as the input fluid. Two different flow were used (0.09 and
1.8 m/day) to test the effect of seepage velocity on geochemical performance. Results of the column tests showed that
heterogeneous flow (preferential flow) conditions developed as a result of mineral precipitation and gas production in the
column. Tracer tests in the columns showed that pH and hydrologic residence time were closely linked. Therefore, pH
may be a key indicator of residence time in PRB installations.
1.1.3
Preliminary results of EPA's long-term performance study were reported in Wilkin et al. (2002). Geochemical and
microbiological factors that control long-term performance of PRBs were evaluated at the Elizabeth City, NO and the
Denver Federal Center, CO sites. These ground-water treatment systems use zero-valent iron granules (Peerless Metal
Powders, Inc.) to intercept and remediate chlorinated volatile organic compounds at the Denver Federal Center (funnel-
and-gate system) and overlapping plumes of hexavalent chromium and chlorinated compounds at Elizabeth City
(continuous wall system). Zero-valent iron at both sites is a long-term sink for C, S, Ca, Si, N, and Mg. Based on an
analysis of mineral precipitate abundance in core materials, after about four years of operation the average rates of
inorganic carbon (1C) and sulfur (S) accumulation were determined to be 0.09 and 0.02 kg/m2y, respectively, in the
Elizabeth City PRB where upgradient waters contain <400 mg/L of total dissolved solids (TDS). At the Denver Federal
Center site, upgradient ground water contains 1000-1200 mg/L TDS and rates of 1C and S accumulation were
determined to be as high as 2.16 and 0.80 kg/m2y, respectively. At both sites, consistent patterns of spatially variable
mineral precipitation and microbial activity were observed. Mineral precipitates and microbial biomass accumulate the
fastest near the upgradient aquifer-Fe° interface. After four years, maximum net reductions in porosity, due to the
-------
accumulation of sulfur and inorganic carbon precipitates, range from 0.032 at Elizabeth City to 0.062 at the Denver
Federal Center (gate 2). While pore space has been lost due the accumulation of authigenic components, neither site
showed evidence of pervasive pore clogging after four years of operation.
The following sections of this report provide descriptions of the Elizabeth City and the Denver Federal Center PRB sites,
and results and analysis of the first five-year monitoring period for contaminant distributions, ground-water chemistry, as
well as mineralogical and microbiological characterization of material that has accumulated within the reactive barriers.
-------
2.0
The U.S. Coast Guard Support Center (USCG-SC) site near Elizabeth City, North Carolina, and the Denver Federal
Center (DFC) in Lakewood, Colorado were evaluated in the EPA portion of the Tri-Agency PRB initiative. Both of these
PRB installations are among the oldest full-scale systems available for study. The Elizabeth City PRB was installed in
June 1996, and the Denver Federal Center system was installed in October 1996. The two sites are of similar age; they
use the same type of iron (Peerless Metal Powders, Inc.), yet they have contrasting ground-water chemistry and different
design configurations (continuous wall configuration at Elizabeth City vs. funnel-and-gate design at the Denver Federal
Center). Table 2.1 provides a general comparison of the PRBs at Elizabeth City and the Denver Federal Center.
2.1 U.S.
The USCG-SC is located about 100 km south of Norfolk, Virginia and 60 km inland from the Outer Banks region of North
Carolina. The base is situated on the southern bank of the Pasquotank River, about 5 km southeast of Elizabeth City,
North Carolina. A hard-chrome plating shop was in operation for more than 30 years in Hangar 79, which is only 60 m
south of the river {Figure 2.1). Following its closure in 1984, soils beneath the shop were found to contain chromium
concentrations up to 14,500 mg/kg. Subsequent site investigations by U.S. EPA personnel identified achromate plume
extending from beneath the shop to the river. The plume has high (>10 mg/L) concentrations of chromate, elevated
sulfate (to 150 mg/L), and minor amounts of volatile chlorinated organic compounds: TCE, cis-DCE, and vinyl chloride
(VC). The plating shop soils and related ground-water contamination are referred to as solid waste management unit
(SWMU) number 9 by the of North Carolina and the USCG. Sampling results from a monitoring network consisting
of more than 40 monitoring wells and about 100 Hydropunch™ and Geoprobe™ monitoring points indicate that the
Cr(VI) plume is about 35 m wide, extends to 6.5 m below ground surface and extends laterally about 60 m from the
hangar to the Pasquotank River (Figure 2.1). Multilevel samplers installed near the barrier wall location indicate that the
bulk of the contamination resides from 4.5 to 6.5 m below ground surface.
The site geology has been described in detail elsewhere (e.g., Puls et a!., 1999b), but essentially consists of typical
Atlantic coastal plain sediments, characterized by variable sequences of surficial sands, silts and clays. In general, the
upper 2 m of the aquifer consists of sandy- to silty-clays that pinch out toward the north, or near the Pasquotank River,
where sandy-fill predominates. Fine-sands, with varying amounts of silt and clay, and silty-clay lenses form the rest of
the shallow aquifer.
Ground-water flow velocity is extremely variable with depth, with a highly conductive layer at roughly 4.5 to 6.5 m below
ground surface. As noted above, this layer coincides with the highest aqueous concentrations of chromate. The ground-
water table ranges from about 1.5 to 2.0 m below ground surface and the average horizontal hydraulic gradient varies
from 0.0011 to 0.0033. Slug tests conducted on monitoring wells with 1.5 m screened Intervals between 3 and 6 m below
ground surface indicate hydraulic conductivity values of between 0.3 to 8.6 m/d. A multiple borehole tracer test in wells
screened between 3.9 to 5.9 m below ground surface was conducted. Ground-water velocities between about 0.13 and
0.18 m/d were measured in this test. Assuming an average hydraulic gradient of 0.0023 and an effective porosity of 0.38,
these flow velocities correspond to an average hydraulic conductivity of about 26 m/d.
In June of 1996, a 46 m long, 7.3 m deep, and 0.6 m wide permeable reactive barrier (continuous wall configuration) of
zero-valent iron (Peerless Metal Powders, Inc.) was installed approximately 30 m from the Pasquotank River (Figure 2.1;
Blowes et al., 1999a,b). The reactive wall was designed to remediate hexavalent chromium-contaminated ground water
and portions of the larger overlapping plume of volatile chlorinated organic compounds. A detailed monitoring network
of over 130 subsurface sampling points was installed in November of 1996 to provide detailed Information on spatial and
temporal changes in pore water geochemistry and hydrology (Blowes et al., 1999a; Puls et al., 1999a).
2.2
The Denver Federal Center (DFC) is located about 10 km west of downtown Denver, Colorado. Aquifer materials at the
site consist of alluvial sediments that overlie the Denver Formation. The Denver Formation is Paleocene to Late
Cretaceous In age and consists of brown, yellowish-brown, gray, and blue-gray Intercalated sandstone, claystone,
siltstone, shale and conglomerate containing olive-brown andesitic sandstone beds. It lies about 2 to 14 m below ground
-------
Table 2.1. Comparison of PRBs Investigated in this Study
Contaminants
PRB Date Iron Iron Ground
Configuration Installed Dimensions Volume water,
SC
U.S.
Coast
Guard
Support
Center
Cr(VI)
TCE, cis-DCE
Continuous
wall
6/96
46 m length
7.3 m deep
0.6 m wide
150mJ
307±149
(n=18)
Ground Ground
water, water,
pH DO
(mg/L)
5.86±0.25
(n=18)
0.7±0.5
(n=17)
Denver
Federal
Center
TCE, TCA, cis-
DCE
Funnel-and-
gate
Gate 1
10/96
12.2m
length
8.5 m deep
1.8 m wide
187mJ
1236±65
(n=3)
7.14±0.15 0.5±0.2
Gate 2
GateS
12.2m
length
9.5 m deep
1.2 m wide
12.2m
length
7.3 m deep
0.6 m wide
139mJ
53
1358±10
(n=3)
1306±10
(n=2)
7.19±0.08 0.2±0.1
7.06±0.07 <0.05
Notes: Geochemical parameters from Elizabeth City are average values (± 1 s.d.) from upgradient monitoring well MW48. All
parameters monitored quarterly from 2/97 to 8/01 and biannually since 8/01. Geochemical parameters from DFC gates 1, 2, 3 are
average values of wells GSA21, GSA 26, and GSA31 from 7/00 to 7/01. SC is specific conductance. DO is dissolved oxygen. DFC
gate 4 was not studied in this investigation.
surface at the DFC and can attain a thickness of up to about 260 m. The Denver Formation has been divided into two
zones, the upper weathered zone and a lower unweathered zone. These two zones are lithologically similar but differ in
color. The upper weathered zone is up to 7 m thick and exhibits a grayish brown color with yellowish orange staining
while the lower unweathered zone has a diagnostic blue color, commonly called "Denver Blue."
There are two separate deposits of alluvial sediments in the vicinity of the DFC. The Verdos Alluvium of Pleistocene age
is a poorly sorted, stratified gravel containing lenses of sand, silt, and clay. In some surface-drainages, the Denver
Formation may be overlain by the Piney Creek Alluvium, which consists of well-stratified sands, silts and clays with
interbedded gravels.
Ground water in the alluvial sediments at the site generally moves from west to east with an average hydraulic velocity
of about 0.3 m per day and a range between about 0,03 and 0.5 m per day {Pacific Western Technologies, 2000).
Shallow ground water Is contaminated with volatile organic compounds including TCE, cis-DCE, VC, 1,1,1 -trichloroethane
(1,1,1-TCA), and 1,1-dichloroethylene (1,1-DCE). At the eastern boundary of the site, maximum concentrations entering
2 of TCE, cis-DCE, 1,1,1 -TCA, and 1,1 -DCE were about 80 u,g/L, 1.6 u,g/L, 200 u,g/L, and 230 u,g/L, respectively,
when the PRB was constructed in November 1996 (FHWA, personal communication, 2002). At one source of
-------
MW35D
PRB
ML15
MW49
CO
r->
CD
z
Q
5
DO
0
MW38
MW Compliance well
© (5 - 10 ft. screen)
ML Multilevel bundle
• (6 in. screens)
MWli
0
10m
Approx, Scale
HANGAR 79
Plating j
Shop |
Ground water
flow direction
Figure 2.1 Plan view map of the PRB at the U.S. Coast Guard Support Center, Elizabeth City, NC.
these volatile organic compounds was a leaking underground storage tank located near Building 52 that was used by the
Federal Highway Administration (FHWA) to store waste, primarily 1,1,1-TCA. Ground-water flow from the aquifer
discharges into Mclntyre Gulch (Figure 2.2). Mclntyre Gulch is a deep channel that penetrates the aquifer along the
southern edge of the contaminant plume. Downing Reservoir is too shallow to be influenced by the aquifer, but the
reservoir stage does affect the ground-water level.
In the fall of 1996, FHWA and the General Services Administration (GSA) installed a permeable reactive barrier at the
eastern edge of the DFC property along north-south trending Kipling Street (Figure 2.2). In contrast to the continuous
wall design used at the USCG-SC, the DFC PRB has a funnel-and-gate design configuration. The funnel component of
the PRB employs metal sheet pile that was driven into unweathered bedrock of the Denver Formation or into resistant,
weathered layers of the Denver Formation. The depth of penetration of the funnel ranged from about 7 to 10 m
(Figure 2.3). The PRB has 4 reactive gates, each 12.2 m long, 9.5 m deep, and from 1.8 m to (gate 1) to 0.6 m (gates
3 and 4) wide (Table 2.1). The design thickness varied because of anticipated differences of contaminant fluxes to the
PRB at different locations. Peagravel zones (0.6 to 1.2 m) were installed immediately upgradient and downgradient of
the reactive iron zones to improve hydraulic connection between the aquifer and the PRB.
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0 m 75
N
i
. r
Figure 2.2 Plan view map of the PRB at the Denver Federal Center, Lakewood, CO (after McMahon et al., 1999).
1700 —
0
183
Distance, m
275
366
Figure 2.3 Schematic cross-section of the Denver Federal Center funnel-and-gate system (after McMahon et al.,
1999).
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3.0
3.1
Ten ground-water monitoring wells (two-inch PVC) located upgradient and downgradient of the iron wall were sampled
on a quarterly basis from February 1997 to August 2001. Beginning in August 2001, sampling of these 10 compliance
wells switched to a biannual schedule. In order to obtain more detailed spatial information on ground-water geochemistry
and contaminant distributions in the subsurface, up to 130 multilevel well bundles were sampled on an annual basis. The
monitoring well network at the Elizabeth City site is described in Blowes et al. (1999b). The locations of the monitoring
wells and the multilevel well bundles in relation to the PRB at the Elizabeth City site are shown in Figure 2.1.
Methods of ground-water and soil core sampling, preservation and analysis, in addition to Quality Assurance/Quality
Control (QA/QC) procedures used in this five-year investigation of PRB performance are described and discussed in
Volume 2 of this EPA Report series.
3. 1. 1
Ground-water contaminants near the PRB location at the Elizabeth City site are hexavalent chromium, TCE, cis-DCE,
and VC. Figure 3.1 shows the time-dependent variations in the concentrations of chromium, TCE, and cis-DCE in
monitoring wells MW48 and MW13, wells located upgradient of the PRB (see Figure 2.1). The concentration trends
evident in these upgradient wells are non-steady state, particularly for TCE in MW48 (Figure 3.1). Concentrations of
TCE in well MW48 have increased with time suggesting that the flux of TCE to the center portion of the reactive barrier
also has increased over the initial five-year period of operation. Average, minimum, and maximum contaminant
concentrations determined at upgradient monitoring points are listed in Table 3.1. The greatest concentrations of
chromium and volatile organic compounds are present in ML21, the multilevel well bundle located just upgradient of and
near the mid-point of reactive barrier. In ML21, the maximum TCE concentration observed was 42,400 jig/L at a depth
of 6.4 m below ground surface (Table 3.1). Note that this concentration is much greater than the average, depth-
integrated concentration of 2,585 ng/L from ML21, and the average TCE concentration in the proximate well MW48
(716 ug/L), which is screened over a 10 ft interval from 4.3 m to 7.3 m below ground surface. The detection of TCE
degradation products, cis-DCE and VC, in upgradient monitoring wells suggests that partial anaerobic bioattenuation of
TCE is occurring in the plume before it reaches the PRB.
The concentration of chromium entering the PRB is greatest near ML21 (average concentration 0.82 mg/L) and least
near ML31 on the west side of the PRB (average concentration 0.032 mg/L). Taking an average chromium concentration
of 0.5 mg/L over the depth interval from 3 m to 6 m below ground surface and an average flow velocity of 0.16 m/d, it is
estimated that the reactive wall removes about 4.1 kg Cr per year. Over the first five years of operation an estimated
21 kg of chromium has been removed from the ground-water plume and sequestered into immobile forms in the solid
phase.
Table 3.2 lists the concentrations of contaminants in monitoring wells located downgradient of the Elizabeth City PRB.
Over the five year period from 1996 to 2001, there were a total of 150 sampling events in downgradient monitoring wells
MW47, MW49, MW50, and in the multilevel well bundles ML15, ML25, and ML35. Chromium was detected in 13
sampling events or in about 9% of the sampling events of wells located in downgradient positions. The highest
concentration of chromium observed in a well downgradient of the PRB from 1996 to 2001 was 0.005 mg/L, significantly
below the MCL and target treatment level for chromium of 0.050 mg/L. Where influent concentrations of chromium were
the greatest, the best performance was observed in that 0 samples of 32 from multilevel well bundle ML25 showed
detectable concentrations of chromium. The monitoring results presented in Table 3.2 show the sustained performance
of the PRB for removing chromium from the ground-water plume. A trend of increasing chromium concentrations at
downgradient locations is not evident after five years of operation.
Average, minimum, and maximum concentrations of TCE, cis-DCE, and VC in downgradient monitoring wells are shown
in Table 3.2. Inspection of the data in Table 3.2 reveals that average TCE concentrations in wells downgradient of the
PRB have been significantly reduced compared to locations upgradient of the PRB, although average concentrations in
downgradient monitoring wells MW49, MW50, and multilevel well bundle ML25 are above the MCL of 5 ug/L. Monitoring
-------
O)
LlT
o
Q
w
'o
60-
50-
40-
30-
20-
10
0
O)
LU
O
6/1/1996 10/14/19972/26/1999 7/10/2000 11/22/2001
1800
1600-
1400-
1200-
1000-
800
600
400
200
0
6/1/1996 10/14/19972/26/1999 7/10/2000 11/22/2001
O)
E
s_T
o
4
3
2
1
0 -
6/1/1996 10/14/19972/26/1999 7/10/2000 11/22/2001
Time
Figure 3.1 Concentrations of contaminants through time in monitoring wells located hydraulically upgradient of the
Elizabeth City PRB (screened interval in MW48 and MW13 is 4.3 to 7.3 m below ground surface).
well MW50, in particular, has typically shown TCE concentrations above 100 u,g/L, especially prior to 2000 (Figure 3.2).
Well MW50 is screened over a 5-foot interval from 7.6 m to 9.1 m below ground surface, a depth interval that extends
below the PRB (maximum depth 7.3 m). Some underflow of the chlorinated solvent plume, therefore, is evident in the
central portion of the PRB. This may have been caused by the disturbance of an unrecognized TCE source near the
PRB and/or a large influx of recharge water into the aquifer near the PRB following installation and prior to repaying of
the parking lot covering the excavation site.
Monitoring results from multilevel bundles ML21 and ML25 indicate that concentrations of cis-DCE have been reduced
from an average of 154 u,g/L in upgradient ground water to an average concentration of 14 u,g/L in downgradient ground
water. Similarly concentrations of VC have been reduced by the PRB. The average VC concentration in multilevel well
bundles ML21 and ML31 was 31 u,g/L and 22 u,g/L, respectively, from 1997 to 2001 (Table 3.1). In downgradient
multilevel well positions ML25 and ML35, VC was detected in about 45% to 55% of the sampling events. When detected
the average VC concentration in multilevel well bundles ML25 and ML35 was 4.3 u,g/L and 2.3 u,g/L, respectively. In
general, concentrations of cis-DCE and VC decrease by about an order of magnitude as a consequence of processes
that take place within the PRB.
10
-------
3.1. Contaminant Concentrations in Ground Water Upgradient from the PRB at Elizabeth City
Well
MW48
MW13
ML11
ML21
ML31
Cr (mg/L)
TCE (ug/L)
cis-DCE (|ig/L)
VC (jig/L)
Cr
TCE
cis-DCE
VC
Cr
TCE
cis-DCE
VC
Cr
TCE
cis-DCE
VC
Cr
TCE
cis-DCE
VC
•"•total
18
17
18
18
18
18
18
18
32
32
32
32
32
32
32
32
30
30
30
30
Selected
18
17
12
3
18
18
6
0
25
24
28
2
25
30
15
12
17
30
16
16
Average Minimum
0.53 0.08
716 111
8.31 <0.1
4.0 <0.1
3.8 2.9
17 <0.1
0.8 <0.1
<0.1
0.75 <0.002
31 <0.1
11 <0.1
1.5 <0.1
0.82 <0.002
2585 <0.1
154 <0.1
31 <0.1
0.03 <0.002
249 2.6
34 <0.1
22 <0.1
Maximum
1.6
1625
60
5.8
4.7
62
6.5
<0.1
2.5
89
56
1.5
3.43
42400
384
52
0.08
1160
103
43
Notes: ntotai is the total number of samples analyzed; ndetected is the number of samples in which the
contaminant was detected. Entries for ML11, ML21, and ML31 (multilevel bundles) are averages of all
sampling depths. MW48 and MW13 data span the time period of November 1996 to February 2002. ML11
and ML31 data were collected in 1997, 1998, and 2000 (annually). ML21 data were collected in 1997, 1998,
1999, 2000, and 2001 (annually). Values below the analytical detection limit were excluded from the
average calculation. Concentration units: Cr (mg/L), TCE (ug/L), cis-DCE (ug/L), VC (ug/L).
A more detailed picture of contaminant behavior can be reached by inspecting 2-dimensional concentration profiles,
constructed by contouring monitoring data from the multilevel well bundles. Figures 3.3 through 3.18 show cross-
sectional profiles of various dissolved solutes and geochemical parameters determined on an annual basis at Elizabeth
City from 1997 to 2001. The locations of transects 1, 2, and 3 are shown in Figure 2.1. For each transect in the cross-
sectional profiles, the position of the PRB is referenced to the position of the furthest upgradient multilevel well bundle
(i.e., ML11, ML21, and ML31). The distribution of subsurface sampling points is shown for transect 3 in the lower left
hand corner of Figure 3.3 to provide an indication of the density of sampling data used to construct the contour diagrams.
Cross-sectional profiles for dissolved chromium are shown in Figure 3.3. Long-term trends indicate that chromium
continues to be removed from the ground-water plume after five years of operation. In general, depth-dependent
concentration profiles have remained consistent with time. However, the dissolved chromium plume appears to be
migrating to shallower depths at a rate of about 15 cm/y in the vicinity of transect 2 (Figure 3.3). In the vicinity of
transect 1 there appears to be no vertical movement of the dissolved chromium plume. The upward movement of the
plume in the vicinity of transect 2 may be linked to the emergence of the deep TCE plume as described below.
More complicated transport behavior is evident for TCE based upon monitoring results from the multilevel well bundles.
There are two separate TCE plumes, one shallow that coincides with the chromium plume (approx. 4 to 6 m below
ground surface). This shallow plume containing TCE (plus cis-DCE and VC) is present in each of the three multilevel well
transects. A second deeper plume containing TCE (minus cis-DCE and VC) appears only upgradient of the PRB in the
vicinity of transect 2 (Figure 3.4). The shallow TCE plume is most concentrated and most shallow near transect 3, unlike
11
-------
Table 3.2. Contaminant Concentrations in Ground Water Downgradient from the PRB at Elizabeth City
Well
MW47
MW49
MW50
ML15
ML25
ML35
Cr (ppm)
TCE (ug/L)
cis-DCE (ug/L)
VC (ug/L)
Cr
TCE
cis-DCE
VC
Cr
TCE
cis-DCE
VC
Cr
TCE
cis-DCE
VC
Cr
TCE
cis-DCE
VC
Cr
TCE
cis-DCE
VC
ritotal
18
18
18
18
18
18
18
18
18
18
18
18
32
32
32
32
32
32
32
32
32
32
32
32
^detected
2
15
17
11
4
11
4
8
4
18
17
14
2
8
9
8
0
15
30
18
1
12
14
14
Average Minimum
0.002 <0.002
6.0 <0.1
7.0 <0.1
5.5 <0.1
0.003 <0.002
9.0 <0.1
1.6 <0.1
5.0 <0.1
0.003 <0.002
204 <0.1
15 <0.1
7 <0.1
0.004 <0.002
1.8 <0.1
1.5 <0.1
2.1 <0.1
<0.002
16 <0.1
14 <0.1
4.3 <0.1
0.005 <0.002
3.6 <0.1
4.3 <0.1
2.3 <0.1
Maximum
0.002
30.2
19.7
10.4
0.004
41
2.6
7.0
0.003
548
35
18
0.005
4.9
1.9
7.8
<0.002
81.6
74.6
10
0.005
9.1
13
4.2
Notes: ntotai is the total number of samples analyzed; Selected is the number of samples in which the
contaminant was detected. Entries for ML15, ML25, and ML35 (multilevel bundles) are averages of all
sampling depths. MW47, MW49, and MW50 data span the time period of November 1996 to February
2002. ML15 and ML35 data were collected in 1997, 1998, and 2000 (annually). ML25 data were collected
in 1997,1998, 1999, 2000, and 2001 (annually). Values below the analytical detection limit were excluded
from the average calculation. Concentration units: Cr (mg/L), TCE (ug/L), cis-DCE (ug/L), VC (ug/L).
chromium, which is most concentrated near transect 2. These relationships suggest that rather than sharing identical
sources, chromium and TCE are more likely being transported in a hydrologically favorable zone. Comparing both TCE
plumes, the highest TCE concentrations are found in the deep plume at a depth range from about 6 to 7 m below ground
surface immediately upgradient of the PRB near transect 2, Over this depth interval, TCE concentrations increased
almost tenfold from 4,320 (ig/L in 1997 to 42,400 (ig/L in 2001. In downgradient multilevel well bundle ML25, TCE
concentrations are below 10 jig/L in all but five samples collected between 1997 and 2001. However, TCE
contamination extends beneath the maximum depth of the PRB near transect 2 and this explains the higher TCE
concentrations detected in MW50 (Figure 3.2).
Unlike TCE, the occurrences of cis-DCE and VC are generally restricted to the shallow plume or to depths of about 4 to
6 m below ground surface (Figures 3.5 and 3.6). In the vicinity near transect 1, concentrations of cis-DCE are reduced
to below detection limits as the contaminant plume passes through the PRB. In this region, the maximum concentration
of cis-DCE observed upgradient of the PRB was 260 jig/L (Figure 3.5). The concentration of cis-DCE in regions
immediately upgradient of the PRB increased from 1997 to 2000. The fact that neither cis-DCE nor VC are present in the
12
-------
LJJ
O
600-
500-
400-
300-
200-
100-
0-
6/1/1996
10/14/1997
2/26/1999
Time
7/10/2000
11/22/2001
Figure 3.2 Concentrations of TCE (u,g/L) through time in monitoring wells located hydraulically downgradient of the
Elizabeth City PRB (screened intervals: MW47, 4.3-7.3 m; MW49, 4.3-7.3 m; MW50, 7.6-9.1 m).
deeper plume near transect 2 suggests that appreciable bioattenaution of TCE in this area has not occurred. In transects
2 and 3, concentrations of cis-DCE above 10 u,g/L are sporadic in downgradient regions; however, with time,
concentrations of cis-DCE appear to increase in regions downgradient of transect 2.
3.1.2 Geochemical Parameters
In this section, results for three geochemical parameters are presented and discussed: pH, Eh, and specific conduc-
tance. Table 3.3 lists average values for these parameters from the upgradient (ML11, ML21, and ML31), downgradient
(ML15, ML25, and ML35), and iron wall (ML14, ML24, and ML34) multilevel samplers. The results of all field analyses
are averaged to provide an overall view of the PRB impact to ground water chemistry. At Elizabeth City, the pH of
ground-water entering the PRB is approximately 6, but varies in time and space from about 5.4 to 6.6. The pH increases
to an average value of 9.5 within the iron media, and then decreases to a value of 7.7 in the aquifer approximately 0.5 m
downgradient of the PRB. This trend in pH clearly shows that the PRB has an impact on the aqueous chemistry of the
downgradient aquifer (Table 3.3). Similarly the Eh of water entering the PRB is moderately oxidizing (approx. 218 mV)
and becomes moderately to highly reducing within the reactive media (approx. -260 mV). Moderately reducing water
emerges from the PRB (approx. -2 mV). The decrease in Eh and increase in pH across the flow path are expected
trends that result from the corrosion of iron in water as discussed in detail in a following section of this report. The
specific conductance of ground water downgradient of the PRB is approximately 15% to 30% lower than that of
upgradient ground water due to partitioning of dissolved solutes into the solid phase within the reactive media.
Interestingly, dissolved oxygen results (data not shown) show essentially no variability between upgradient, iron, and
downgradient positions. The reasons for this are likely due to the difficulty in obtaining highly accurate DO
concentrations at low levels and in the presence of ferrous iron.
Long-term trends in pH, Eh, and specific conductance are shown in Figures 3.7-3.9. In transect 2, the high pH zone
caused by corrosion of the zero-valent iron has remained largely unchanged in space over the initial five-year period of
operation (Figure 3.7). This result indicates that the zero-valent iron at Elizabeth City still retains some degree of
reactivity even after five years of subsurface exposure. The pH values measured in upgradient regions have remained
13
-------
-2
1997
1998
2000
01234 01
-3
-7
2.0
1.8
1.6
1.4
1.2
1.0
0.8
0.6
0.4
0.2
2001
0123
0 1
0123
-2
-2
-a
-4
-5
-&
1234 01234
Distance, m
rn i \\ _[\ r^T
flow
0
0
0
0
0
0
0
0
"c
'o
Q_
D)
C
"5.
E
CD
-------
-2
1998
2000
0123401234 012
-7-
'6500
950
8000 — 1100
"- OJ CM oo • 800
I"" 23^i500°1 650
20 Ł 3500Ł 500
15 S 2000 2 350
200
50
-100
|-To°Oo|
-2
401234 01234
Distance, m
flow
Q
D
ED -|
H '§.
Q ?
Q t
Q CO
D
i i rn
QQQ
Q
Q
D
HBQ
Q
Q Q
Q
D D
Q
D D
DQD
QD
3
D
D
D
D
D
D
Q
D
01234
Figure 3.4 Cross-sectional profiles showing TCE concentrations (|^g/L) in transects 1, 2, and 3, Elizabeth City PRB.
15
-------
-2
1997
-2
1998
2000
Transect 1
012
!2 26 3C
6 1014182
Transect 2
-50 10 70 130 190 250
Transect 3
-5 5 15 25 35 45 55
c/s-DCE, |Lig/L
2001
123 0123
~TZ] UIIJ CT~
flow
H
H
H
H
H
H
H
H
(/)
'c
'o
Q.
O)
.^
Q.
CD
H
Q Q
Q
Q
Q
Q
Q Q
Q
Q Q
Q
Q Q
Q
Q Q
Q
Q Q
Q
n
Q
Q
Q
Q
Q
Q
H
Q
01234
Distance, m
Figure 3.5 Cross-sectional profiles showing cis-DCE concentrations (|JXj/L) in transects 1, 2, and 3, Elizabeth City
PRB.
16
-------
1997
1998
2000
I''
ho
8.8
36
C\| CO
7.1 -^—29 -H _ 23
5.5 w
c
-7
01
34 01
IE I
CO
I29 0|
I
16 2
Iz.j _ 8.8
0.63 I 1.9
-1.0 J-5.0
VC,jag/L
40
34
29
3.9 ro
2.3
2001
0 1
0123401234 01234
Distance, m
-2 =
-3
-4
-5
-6
-7
i i
a
a
a
a
a
a
a
H
flow
i
u)
'c
I
O)
.^
Q.
m
OT
i i i i
i i
[7]
a a
a
[T]
a
a
Q Q
a
a a
GH
a a
[T]
a a
a
a a
[T]
a
a
a
a
a
a
a
a
a
01234
18
12
6.3
0.63
-5.0
Figure 3.6 Cross-sectional profiles showing VC concentrations (|JXj/L) in transects 1, 2, and 3, Elizabeth City PRB.
17
-------
Table 3.3. Geochemical Parameters in Upgradient, Iron Wall, and Downgradient Locations, Elizabeth City PRB
Location
Upgradient
Iron wall
Downgradient
Eh
(mV)
218
(±311)
-258
(±212)
-2
(±190)
PH
6.02
(± 0.24)
9.53
(± 0.72)
7.71
(±1.14)
Specific Conductance
((aS/cm)
390
(±177)
186
(±105)
192
(±110)
Notes: Average value and 1 s.d. reported for upgradient (ML11, ML21, and ML31), iron wall (ML14, ML24,
and ML34), and downgradient (ML15, ML25, and ML35) well locations.
largely unchanged in time and space. Similarly, temporal and spatial distribution of pH within the iron media and in
downgradient regions has remained consistent since 1997 (Figure 3.7). Transects 1 and 3 also show fairly consistent
patterns in pH. Notable is a relatively narrow zone near the influent side of the PRB where increases of 3 to 4 pH units
are observed; a sharp gradient in pH has been observed in each sampling event since 1997.
Iron corrosion is expected to result in moderately high pH and low Eh conditions. The contouring results shown in
Figures 3.7 and 3.8 suggest, however, that pH and Eh values are not directly coupled (Wilkin, 2002). Values of Eh in the
iron are gradually increasing with time (Figure 3.8), whereas pH has been consistent with time. For example, in transect
2, the cross-sectional area represented by Eh values below -100 mV has progressively decreased from 1997 to 2001
(Figure 3.8). There still exists after five years a sharp gradient in redox between the relatively oxidizing upgradient
aquifer and the moderately to highly reducing reactive zone. The redox potential of upgradient waters is as high as +500
mV and values less than 0 mV are typical within the reactive media. After the first five years of operation, the redox
gradient continues to support efficient removal of hexavalent chromium and chlorinated compounds. Nevertheless, a
trend of decreasing reductive capacity with time is evident.
Corrosion of iron metal results in a moderately alkaline pH and low Eh geochemical environment that drives abiotic
mineral precipitation and supports a variety of microbial metabolic pathways. The net effect of mineral precipitation is
broadly captured in measurements of specific conductance (SC) that reflect an overall decrease in the concentration of
total dissolved solids through the reactive barrier. In transect 2, the highest concentrations of ground-water solutes are
observed at a depth of about 4 to 6 m below ground surface (Figure 3.9). This depth interval also corresponds to a
maximum in SC and a zone within the PRB in which core analyses show enrichments in solid-phase concentrations of
inorganic carbon, sulfur, and microbial biomass. Over the initial five-year period, variability in SC over a factor of about
2 has been observed in upgradient regions at the Elizabeth City site. Variability of specific conductance in upgradient
regions is also reflected in the temporal and spatial distribution of SC values within the reactive media and in
downgradient regions (Figure 3.9). The consistent trend of decreasing SC values through the reactive media suggests
that mineral precipitation processes are operating over the lifetime of the reactive media, i.e., that the zero-valent iron has
remained reactive over the first five years of operation. This sustained reactivity is likely due to the presence of the steep
pH and/or Eh gradient as described above. The region where a minimum in Eh (most reducing) is observed corresponds
to the depth of the SC maximum in upgradient ground water. This observation suggests that the concentration of
ground-water solutes, or more likely the concentration of terminal electron acceptors such as sulfate and microbial
activity, is perhaps what leads to the Eh minima rather than only anaerobic iron corrosion reactions. Furthermore, the
subtle changes in Eh within the reactive media also seem to be time-dependent and reflect temporal changes in
upgradient Eh and SC.
3.1.3
The dominant cations in ground water at the Elizabeth City site are sodium, potassium, calcium, and magnesium. On a
molar basis, sodium is the dominant cation followed by Ca>Mg>K. From 1997 to 2001, upgradient ground water
contained up to 5.2 mM Na, 1.1 mM Ca, 0.8 mM Mg, and 0.2 mM K. Ground water most concentrated in dissolved
solutes was consistently found at 4 to 6 meters below ground surface in each of the multilevel well bundles (Figure 3.9).
Inspection of cross-sectional profiles in Figures 3.10-3.12 indicates that the zero-valent iron effectively removes calcium,
magnesium, and to some extent, sodium, from ground water at the Elizabeth City site.
18
-------
-2
1997
1998
-2,
-2T
2000
2001
123 0123
0123
-2-
234
Distance, m
E
E
E
El
E
E
E
E
flow
«
'o
Q.
g>
"5.
CO
C/3
1 — 1 1 — 1
B
B B
E
E
B
B
B B
E
E E
B "B
E
E E
B
B B
E
B
e..
B
B
E
B
E
E
B
E
01234
Figure 3.7 Cross-sectional profiles showing pH distributions in transects 1, 2, and 3, Elizabeth City PRB.
19
-------
1997
1998
2000
-7-
J
6 1 2 3 4
-4
-5
-6
-7
0 1
-2
-3-
-4-
-&
-6-
-7-
-7J
1 2 3
-4-
-S
-&
-7-
/ \
01
-7-
0123401234 01234
Distance, m
> ^
-300 Ł LL?
-500
LJJ
2001
C/)
-4
-5
-6
-7
(
1
/ \
y
\
\
, \
\ \
/ / \
l
J j
\ / /
/
/
/
/
\
\ \ \
1
\
/
/I
/I
) 123
i—
G
G
Q
G
G
G
G
Q
flow
S
'o
Q.
D)
"5.
E
CD
n i — i
LP
Q 'Q
Q
G
G
Q "a
a "a
G
Q a
Q^Q
Q
a a
"a
a
a
a
a
Q
a
a
a
01234
Figure 3.8 Cross-sectional profiles showing Eh distributions (mV) in transects 1, 2, and 3, Elizabeth City PRB.
20
-------
1997
1998
-2-,
2000
01234 01234
1234
800
700
600 Q
500 ^
400 .2
300 Ł "
200 O "O
100 Q_ O
0
2001
o
0123
-1-
0123
1 2 3
-1- x -7-N. / -7J
0 1 2 3 4 0 1 2 3 4 01234
Distance, m
D
Q
D
Q
D
D
Q
D
6
flow
*2
c
o
CL
D)
Ł=
^
E
CD
CO
1
— rj — i i — i
UL
[
D
m
t
[
D
0
[
D
[
0
[
Q
[
V
]
D
]
D
]
D
H
]
D
]
H
]
D
D
c^
D
D
D
m
D
m
D
D
234
Figure 3.9 Cross-sectional profiles showing specific conductance distributions (|o,S/cm) in transects 1, 2, and 3,
Elizabeth City PRB.
21
-------
-2
1997
0 1
1998
2000
50
45
40
35
30
25
20
15
10
5
0
O)
E
co
O
2001
0123
1 2 3
0123
0123
-2
LJ
E
E
E
E
B
B
D
B
co
'o
CL
__
c
"o_
E
CD
W
UL^
B
E B
E
B
B
B
B B
B
B B
B
B B
D
B B
B
B B
B
B
B
B
B
B
B
B
B
B
01234
Distance, m
Figure 3.10 Cross-sectional profiles showing calcium concentrations (mg/L) in transects 1, 2, and 3, Elizabeth City
PRB.
22
-------
-2
1997
1998
2000
-7.
01234
1234 01234
-4
-5
-6
-7
0123
01 23
1 23
16
13
10
7
4
1
-2
O)
E
2001
01 23
-2-
-2T
01234 01
Distance, m
— c u^j —
0
0 0 0
0
0 "a
0«2 a^a
D •§ D^D
Q. Q
DD> a a
~ D
E" Q
D CD D D
a 'H
— r^r
0
a
0
D
0
0
0
0
01234
Figure 3.11 Cross-sectional profiles showing magnesium concentrations (mg/L) in transects 1, 2, and 3, Elizabeth
CityPRB
23
-------
1997
1998
2000
110
90
70
50
30
10
r
-10 03
O)
E
0123401234 01234
-3
CD
CO
«,*
-6-
-7
_ \
-3
-6-
-7-
-4
-5
-&
-7
0123 0123
-3
2001
-4
-S
-7-
0123 0123
0)
CO
03 CD .5
Q
-3
-4
-5
-6
-7
A
\ 30
56 \
/
^30^\
^-^
^X^
-2
-3
-4-
-5
-6-
-7-
-2
-3
-4
-5
-6
-7
0123401234 01234
Distance, m
-3
-4-
-5
-6
-7
I-1 L- 'LI-1
0 flow 0 Q
0
0 [D
0 "^ 0 Q
° EL
0 ^ 0 Q
05 0_
0 ^ 0 Q
Q. r-|
0 ^ 0 Q
0^0%
0 Q
cr
0
0
0
0
0
0
0
0
01234
Figure 3.12 Cross-sectional profiles showing sodium concentrations (mg/L) in transects 1, 2, and 3, Elizabeth City
PRB.
24
-------
Concentrations of potassium appear to be largely unaffected by the reactive media (Figure 3.13). In transect 2, calcium
concentrations range from about 5 to 30 mg/L in upgradient ground water and are typically less than 5 mg/L in
downgradient ground water. Depth-dependent removal efficiencies of the major cations are presented in Table 3.4. At
discrete depth intervals, up to 71 % Ca, 73% Mg, and 60% Na are removed as ground water passes through the reactive
media. Calcium and magnesium are consistently removed through time, although percent removal values are lower in
2001 compared to the average value from 1997 to 2001 (Table 3.4). This trend may suggest decreasing reactivity over
time with respect to removal of divalent cations. Sodium shows removal at some intervals and relative gains in
concentrations at deeper levels (Figure 3.12; Table 3.4). The 2-dimensionaI trends in sodium and chloride concentra-
tions suggest that there may be some component of vertical redistribution of solutes within the PRB.
Anions present in ground water at the Elizabeth City site include chloride, sulfate, bicarbonate, and nitrate. On a molar
basis, chloride and bicarbonate are the dominant anions followed by sulfate. Nitrate is commonly detected in upgradient
ground water but always at concentrations below about 5 mg/L. In downgradient multilevel well bundles, nitrate is
typically below the analytical detection limit (0.10 mg/L). In transect 2, sulfate concentrations typically range from about
25 to 90 mg/L in upgradient ground water, but then decrease to less than 1 mg/L in most downgradient multi-level wells.
Inspection of the cross-sectional profiles in Figures 3.14-3.17 indicates that anionic species, especially sulfate and
bicarbonate, are effectively removed by zero-valent iron. Depth-dependent removal efficiencies of the major anions are
presented in Table 3.5. Typically >90% of influent sulfate is removed as ground water passes through the reactive
media. The high removal efficiency of sulfate does not appear to be decreasing with time (Table 3.5). Nitrate is also
effectively removed by the PRB (Table 3.5), although the anomalous trends apparent in 2000 are not understood.
Chloride removal occurs at depths above 5 m but reactions in the PRB appear to be sources of chloride at depths below
5 m. The of chloride is likely related to the degradation of TCE at depths below 5 m.
Silica in upgradient ground water is present at concentrations up to about 16 mg/L, dominantly as the uncharged form
H4SiO4° (aq). Silica concentrations in the reactive media are typically <1 mg/L and levels then rebound in downgradient
wells to values from about 1 to 8 mg/L (Figure 3.18). Similar behavior for dissolved silica was also reported at the Moffet
Field and Lowry Air Force PRBs (Gavaskar et al., 2002).
Over the first five years of operation, the Elizabeth City PRB has consistently removed C, S, Ca, Mg, Si, and N from
ground water. Over this time period there appears to be a slight to no discernible loss in the capacity or efficiency of
removal of these inorganic components. These elements are removed from ground water by mineral precipitation or by
adsorption, processes examined in more detail in later sections of this report.
3.2 at
Core samples from the PRB were collected on an annual basis at Elizabeth City to assess the extent of corrosion and
mineral buildup on the iron surfaces. Core collection methods and analysis procedures are described in Volume 2 of this
EPA Report series. In all cases, 5 cm inner diameter cores were collected using a Geoprobe™. Core barrels were
driven using a pneumatic hammer to the desired sampling location and continuous, up to 110 cm, sections of iron, iron
+ soil, or soil were retrieved. Angle cores (30° relative to vertical) and vertical cores were collected in order the
spatial distribution of mineral/biomass buildup in the reactive media. Prior to pushing the core barrel, an electrical
conductivity profile was collected to verify the exact position of the iron/aquifer interface. Details of core collection
procedures are described in Volume 2 of this EPA Report series. In all cases, core recovery was 50 to 92% of the
expected value. Core materials from the Elizabeth City PRB were jet black in color without any obvious signs of
cementation or oxidation. Immediately after collection, the cores were frozen and shipped back to the Ground Water and
Ecosystems Restoration Division in Ada, OK 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 4 sub-samples: (1)
inorganic carbon analyses, (2) sulfur analyses/X-ray diffraction (XRD), (3) Scanning electron microscopy (SEM)/X-ray
photoelectron spectroscopy (XPS) analyses, and, (4) microbial assays (phospholipid fatty acids, PLFA). All sub-
samples were retained in airtight vials to prevent any air oxidation of redox-sensitive constituents. Details of analytical
methods and QA/QC procedures used to characterize the core materials are presented in Volume 2 of this EPA Report
series.
Locations of coring events at Elizabeth City are shown in Figure 3.19, and information on core recovery, core length, and
depth of core penetration is presented in Table 3.6. Cores collected from the upgradient aquifer/iron region generally
penetrated the PRB at depths of 4.5 to 6.5 m below ground surface. As such, the cores were retrieved from portions of
the PRB where the highest concentrations of chromium and total dissolved solutes entered the reactive media (see, e.g.,
Figure 3.3).
3.2.1
Based on long-term trends in ground-water concentrations of bicarbonate and sulfate, significant accumulations of
inorganic carbon and sulfur precipitates might be expected in the PRB at Elizabeth City. In order to confirm this
25
-------
-2T
1997
-2
1998
01234 01
2000
34 01234
10.0
8.0
6.0
4.0 -->
D)
2.0 Ł
0 k 7
2001
-4
-5
0123
-2
-2
-4
-5
-&
-2
-4
-5
-&
-7
1234
-B-
'flow
Q H
Q^D Q
a. Q
Q E
CD
Q OT
n
a a
Q
Distance, m
01234
Figure 3.13 Cross-sectional profiles showing potassium concentrations (mg/L) in transects 1, 2, and 3, Elizabeth City
PRB.
26
-------
Table 3.4. Changes in Concentration of Metals in Ground Water at Transect 2 (Elizabeth City) as a Function of
Time and Depth
Average values from 1997 to
Depth,
m
4
5
6
7
4
5
6
7
4
5
6
7
4
5
6
7
ML21
Upgradient
Ca, mg/L
26.6
8.6
10.3
13.7
Mg, mg/L
12.0
6.3
6.5
8.1
Na, mg/L
52.6
66.3
27.6
23.0
K, mg/L
4.8
1.7
1.3
1.1
ML25
Downgradient
8.1
4.8
4.5
4.0
3.2
3.5
2.6
2.3
20.9
47.1
42.1
34.1
3.5
2.4
2.3
0.7
2001
%
Change
69.5
44.2
56.3
70.8
73.3
44.4
60.0
71.6
60.3
29.0
-52.5
-48.3
27.1
-41.2
-76.9
36.4
Values from
ML21
Upgradient
18.8
7.4
10.1
12.7
10.0
5.9
6.8
7.9
78.8
35.6
20.7
23.2
4.4
3.6
1.8
1.3
2001
ML25
Downgradient
12.4
4.3
5.9
6.2
5.2
5.1
4.3
3.5
19.8
34.5
33.5
24.4
3.9
3.1
2.1
1.1
o/
/o
Change
34.0
41.9
41.6
51.2
48.0
13.6
36.8
55.7
74.9
3.1
-61.8
-5.2
11.4
13.9
16.7
15.4
% change calculated using:
[(upgradient concentration-downgradient concentration)/(upgradient concentration)]x100
expectation and to document the concentration distribution of inorganic precipitates within the reactive media, solid
phase analyses of carbon and sulfur were performed. Results of these analyses on Elizabeth City cores are listed in
Table A1 (Appendix 1). Inorganic carbon results are given in weight percent C based upon carbon that is 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.xH2O).
The mineral composition of authigenic precipitates formed within the reactive barrier is more fully explored in the sections
below on X-ray diffraction analysis and SEM analysis. The bulk carbon concentrations are used to access the space-
and time-dependent quantity of inorganic carbon that has deposited in the reactive media, an understanding of which is
necessary for estimating the extent of pore infilling through time.
Results of all carbon analyses on Elizabeth City core materials (n = 170) are presented in Table A1. The inorganic, acid-
extractable carbon concentration in unreacted Peerless iron was found to be 15 |ig/g; therefore, concentrations of
inorganic carbon in core materials above this value are the result of mineral precipitation processes that have taken place
within the reactive media as a result of continued ground-water exposure. Concentrations of inorganic carbon within
>90% iron samples range from <1 to 5870 ng/g. In all cases, the highest concentrations were determined in samples
collected adjacent to the upgradient aquifer/iron interface, and the lowest concentrations were detected near the
downgradient edge of the reactive media. Accumulation of inorganic carbon precipitates does not occur at the same rate
throughout the reactive media, rather carbon accumulation is highly spatially variable and is largely restricted to the
upgradient portion of the reactive media (Wilkin et al., 2003).
27
-------
-2-
1997
1998
2000
0123
105
90
75
60
45
30
15
0
E
O
2001
0123
0123
23
Distance, m
0
Qr0
H
flow
0
0
0
0
0
0
0
0
6
CO
"c
a
c
E
CD
C/>
1
0%
0
%
0D0
0^:
0^0
0^:
0D0
^
2
0
0
0
0
0
0
0
0
3 4
Figure 3.14 Cross-sectional profiles showing chloride concentrations (mg/L) in transects 1, 2, and 3, Elizabeth City
PRB.
28
-------
1997
1998
2000
105
-2T
-7
-7
0123401234 01234
Q DEP
flow
Q D D
Q
Q U
Q .1 Q Q
E Q- DTD
D) LL
Q .E LI El
~ EL
D Ł E E
[C 1 1
Li co Q Q
D D
0
D
D
D
D
E
E
E
D
01234
Figure 3.15 Cross-sectional profiles showing sulfate concentrations (mg/L) in transects 1, 2, and 3, Elizabeth City
PRB.
29
-------
1997
-2
-a
-4
-5
-6
-7
1998
2000
0123401234 01
90 Ł
130
120
105 Q)
90
75
60
45
30
15
0
03
<
2001
0123 0123
0123
0123
-2
-2
-4
-5
-6
0123401234 01234
Distance, m
Q
QrO
H
flow
Q
Q
Q
Q
Q
Q
Q
Q
Eft
Q
-2 n
•- Q Q
ra Q^
= Q Q
i" Q%
CD m
w Em
%
Q
H
Q
D
Q
Q
Q
Q
Figure 3.16 Cross-sectional profiles showing alkalinity distributions (mg/L) in transects 1, 2, and 3, Elizabeth City
PRB.
30
-------
-2
1997
1998
2000
I 2.5
'2.0
1.1
0.70
0.25
-0.20
D)
01234 01234
2001
-3
-4
-5
-6
-7
^^~
H
H
H
Q
Q
D
Q
D
flow
C/)
~Ł
'o
Q.
D)
^
"5.
E
CD
C/)
UL^
Q
H H
H
H
H
H
H H
H
Q Q
Q "Q
Q
Q Q
QQQ
Q
D
— pr
Q
H
H
a
H
H
Q
H
123401234 01234
01234
Distance, m
Figure 3.17 Cross-sectional profiles showing nitrate concentrations (mg/L) in transects 1, 2, and 3, Elizabeth City
PRB.
31
-------
Table 3.5. Changes in Concentration of Anions in Ground Water at Transect 2 (Elizabeth City) as a Function of
Time and Depth
Average values from 1997 to 2001
Depth,
m
4
5
6
7
4
5
6
7
4
5
6
7
4
5
6
7
ML21
Upgradient
chloride,
mg/L
66.3
53.5
17.1
21.4
sulfate,
mg/L
47.9
47.8
27.4
16.2
alkalinity,
mg/L
85
54
50
55
nitrate,
mg/L
4.7
1.4
0.5
1.2
ML25
Downgradient
19.5
39.5
39.2
38.3
7.4
0.32
0.48
0.48
56
64
58
33
0.08
0.08
0.5
0.4
%
Change
70.6
26.2
-129.2
-78.9
84.6
99.3
97.8
97.0
34.1
-18.5
-16.0
40.0
98.3
94.3
0
36.1
Values from
ML21
Upgradient
79.7
22.5
13.8
27.1
79.7
32.2
27.9
23.4
65
57
54
54
1.4
0.3
0.4
1.6
2001
ML25
Downgradient
13.9
25.7
38.7
40.5
10.1
<0.1
<0.1
<0.1
79
72
56
28
<0.1
<0.1
<0.1
<0.1
o/
/o
Change
82.6
-14.2
-180.4
-49.4
87.3
>99.7
>99.6
>99.6
-21.5
-26.3
-3.7
48.1
>92.9
>66.7
>75.0
>93.8
% change calculated using:
[(upgradient concentration-downgradient concentration)/(upgradient concentration)]x100
A cross-sectional profile showing the concentration distribution of inorganic carbon in the solid-phase is shown in
Figure 3.20. The concentration profile was constructed based upon three angle cores collected in May 2001 that
intercepted the upgradient edge of the PRB at depths from about 4.8 to 6.2 m below ground surface. The sampling
transect was located approximately 3 m west of multilevel well transect 1 (see Figure 3.19). At the time of sampling, the
highest concentrations of inorganic carbon were found at a depth of about 5 m below ground surface and 5-10 cm inside
of the reactive media/aquifer interface. The greatest amount of carbon accumulation is localized in a rather narrow depth
interval that corresponds to the depth where a maximum in total dissolved solids and bicarbonate enters the reactive
media (compare Figure 3.20 with Figures 3.9 and 3.16). Not surprisingly, "hotspots" of mineral precipitation within the
reactive media are tied to where influent ground water is enriched in dissolved solutes.
The time-dependent accumulation of mineral precipitates is difficult to evaluate, in part, because the core sampling
location, relative to ground-water solute inputs, is obviously important in governing accumulation rates. Unless core
sampling is carried out at the exact same location through time, it may be difficult to evaluate time-resolved data sets.
Figure 3.21 shows inorganic carbon concentrations in angle cores collected in 1997,1998, 1999, 2000, and 2001. Each
of these cores was collected from the same general vicinity so that the trends observed are considered to be
comparable. Shown in Figure 3.21 is an overall regular increase in inorganic carbon concentrations with time. The
trends clearly indicate that a front of precipitation is progressively passing through the reactive media with time. This
observation is critically important because it suggests that complete, rapid pore-infilling does not occur in the region
immediately adjacent to the upgradient aquifer/iron interface. Rather, mineral precipitation occurs in a larger volume of
the reactive media than is present immediately adjacent to the upgradient interface.
32
-------
ML10
(June 2000)
ML20
(June 2000)
ML30
(June 2000)
1234
Distance, m
1 2 3
Distance, m
Distance, m
0 2.0 4.0 6.0 8.0 10 12 14 16
Si, mg/L
Figure 3.18 Cross-sectional profiles showing silica concentrations (mg/L) in transect 2, Elizabeth City PRB.
3.2.2 Sulfur Analysis
Total sulfur measurements were made with a UIC sulfur coulometer system. Iron samples were covered with V2O5 and
combusted in the presence of oxygen at 1050 °C. Evolved gases are passed through a column of reduced Cu to
quantitatively convert all sulfur to SO2, which is then carried to the coulometer cell where it is absorbed and
coulometrically titrated. Results of total sulfur analyses on Elizabeth City cores are listed in Table A1. Un-reacted
Peerless iron contains about 5 jig/g of sulfur using this combustion method. In addition to total sulfur measurements, the
concentrations of acid-volatile sulfide (AVS) and chromium-reducible sulfur (CRS) were determined by chemical
extraction with hot, 6 M HCI and 1 M CrCI2 in 0.5 M HCI, respectively (Zhabina and Volkov, 1978). These acid extraction
methods determine the quantities of metal monosulfide precipitates (AVS) and iron disulfide precipitates (i.e., pyrite;
CRS). Metal monosulfide precipitates, such as disordered mackinawite (Fe1+xS) and mackinawite (Fe1+xS), dissolve in
dilute hydrochloric acid and subsequently release hydrogen sulfide gas which can be measured. Iron disulfide minerals,
pyrite (cubic FeS2) and marcasite (orthorhombic FeS2), and elemental sulfur are not dissolved in dilute hydrochloric acid
but are solubilized by using the more aggressive CrCI2-HCI solution. The hydrogen sulfide gas released in this extraction
step (CRS) can again be quantified and related back to the concentration of disulfide-sulfur present in the core materials.
33
-------
MW49
EC050801-4 o
EC050801-3o
EC060300-4°0
EC050801-1
EC90902
OQQ EC909Q3'
EC6114
EC050801-6
,o ECQ5Q8Q1-7
°EC050801-5
MW48
EC060200-1 O
OEC030616
« ML11 ffi
ML13
Reactive Barrier ML14
--*
--- •
— ML12 ML22 5
EC060300-5 ML23.5-
° 0 ML24.5 —
» ML21
-•
» ML31
•~
*--
ML32 OECQ6Q5QO7
~— ML33
- ML34
EC050901-8,9
• ML25
EC060300-6 O
EC6101 O
ML35 I
°EC010618
3 m
® Compliance well location
• Cluster well location
° Angle or vertical coring
position
N
Figure 3.19. Coring locations and monitoring well locations at the Elizabeth City Permeable Reactive Barrier site (plan view).
-------
3.6
Cores Collected for Analysis at the Elizabeth City PRB
w
01
Sample ID
Location
Date
Angle
Core length
Recovery
Notes: * Core EC6101 captures the upgradient iron/aquifer interface but was collected by pushing the core barrel from the
downgradient side of the reactive barrier. Depth to iron is the depth below ground surface in meters where zero-valent iron was intercepted.
Depth to Fe
EC050801-1
EC050801-3
EC050801-4
EC050801-5
EC050801-6
EC050801-7
EC050901-8
EC050901-9
EC060200-1
EC060300-4
EC060300-5
EC060300-6
EC060500-7
EC010618
EC030616
EC90902
EC90903
EC6101
EC6114
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Vertical
Vertical
Upgradient
Upgradient
Vertical
Downgradient
Vertical
Downgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
June-00
June-00
June-00
June-00
June-00
June-99
June-99
September-98
September-98
June-98
June-98
30
30
30
30
30
30
90
90
30
30
90
30
90
30
30
30
30
30
30
107
99
109
94
102
109
102
112
61
76
84
94
69
61
61
61
61
69
66
88
81
90
77
83
90
83
92
50
63
69
77
56
50
50
50
50
56
54
4.4
5.3
5.8
4.9
6.3
5.6
2.2
2.2
4.8
4.8
2.2
4.0
2.3
5.2
5.5
4.8
5.0
4.8
5.0
-------
Inorganic Carbon
4816 ppm
Q.
CD
Q
8
-0.6 -0.2 0 0.2 0.6
Position, m
<100 ppm
100-1000 ppm
1000-3000 ppm
3000-4000 ppm
>4000 ppm
B
Figure 3.20 Cross-sectional profile showing concentration distribution of inorganic carbon in the solid phase
(u,g/g=ppm), Elizabeth City PRB (June 2002).
36
-------
6000-
5000-
. 4000-
_ *
o
O 3000-
"E
o
c
P> 2000-
1000-
0-
2001
2000
1999
1998a
1998b
0
10
20
30
40
50
Distance, cm
Figure 3.21 Concentrations of inorganic carbon (u,g/g) in core materials through time (Elizabeth City PRB).
Concentrations of total sulfur in the iron media range from about 1 to 3880 u,g/g. The concentration distribution of sulfur
broadly compares with that of inorganic carbon (Figure 3.22). The highest sulfur concentrations are found near the
upgradient interface in regions proximal to where maximum ground-water concentrations of sulfate enter the reactive
zone (see Figure 3.15). The lowest concentrations of sulfur in the solid-phase are found near the downgradient edge of
the PRB.
Sulfur may be present in the core materials as sulfide (disordered mackinawite, mackinawite, greigite), disulfide (pyrite
or marcasite), elemental sulfur, or as sulfate (e.g., sulfate green rust). Total sulfur concentrations will reflect the sum of
all of these sulfur forms. In order to more accurately determine sulfur partitioning and mineralogy, sequential extraction
procedures were carried out (Wilkin et al., 2003). Results presented in Table A2 indicate that over 90% of the total sulfur
is present as sulfide. The remaining sulfur would appear to be made up of iron disulfides (pyrite) and perhaps more acid-
resistant forms of iron monosulfide such as greigite (Fe3S4). Both of these phases have been detected in minute
quantities using high-resolution transmission electron microscopy (Furukawa et al., 2002). Sulfate green rust does not
appear to be an important mineral form in the reactive media at Elizabeth City, which is consistent with geochemical
modeling results. Based on the HCO3VSO42" ratios of ground water at Elizabeth City the carbonate form of green rust is
expected to dominate over the sulfate form (Wilkin et al., 2002). The solid-phase dominance of sulfide over sulfate in the
reactive media and the complete loss of dissolved sulfate demonstrate that reduction of sulfate to sulfide has gone
essentially to completion.
The masses of 1C and S deposited within the iron barriers determined from solid-phase characterization agree
reasonably well with estimated masses based on changes in concentrations of dissolved solutes and flow rates. Mass
balance estimates are presented and discussed in Wilkin et al. (2002). Inorganic carbon mass balance agrees to within
37
-------
Sulfur
Q.
CD
Q
2680 ppm
-0.6 -0.2 0 0.2
Position, m
<100 ppm
100-500 ppm
500-1000 ppm
1000-2000 ppm
>2000 ppm
Figure 3.22 Cross-sectional profile showing concentration distribution of sulfur in the solid phase (u,g/g=ppm),
Elizabeth City PRB (June 2002).
38
-------
a factor of 1.5x for the Elizabeth City PRB, More sulfur accumulation, however, would be expected in the Elizabeth City
PRB, based on changes in sulfate concentrations than has been observed on the core materials. Several factors lead
to uncertainty in mass balance calculations for PRBs. Estimates of mineral accumulation based on changes in ground-
water chemistry critically depend on chemical and especially hydrogeologic measurements. Determination of dissolved
constituents may be analyzed at a high level of accuracy and precision, <5%. Similarly, estimates of mass accumulation
based on characterization of core materials depend on the accuracy of analytical measurements, in addition to estimates
of emplaced iron density. However, estimates of ground water flow volumes moving through PRBs are prone to large
uncertainties, and these dominate the total error of mass balance calculations (20-50%). Spatial variability in ground
water flow velocity, concentration of solutes, concentration of solid phase products, and emplaced iron density all factor
into the uncertainty analysis of mass balance calculations.
Cr
To understand the distribution of solid-phase chromium, dilute hydrochloric acid (1 M) leaches were performed on
upgradient aquifer and reactive iron material. Analysis of un-reacted Peerless iron indicated acid-extractible chromium
concentrations of about 8.8 u,g/g. Concentrations of acid-extractible chromium as high as 72 jig/g were measured in the
reacted iron media after five years of exposure. Chromium is enriched in the solid-phase in subsurface regions near the
upgradient iron/aquifer interface.
3.2.4 X-ray
Powder X-ray diffraction scans for core samples collected in 2000 and 2001 are shown in Figure 3.23. 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-micron filter paper (polycarbonate). The separated particles were mounted on
a zero-background quartz plate and scanned with Cu Koc radiation from 3° to 80° 2-theta using a Rigaku Miniflex
Diffractometer. Diffraction analysis is used to determine the bulk mineralogical composition of the materials removed
from the reactive media. The results can be used to qualitatively evaluate the abundance of various mineral phases; no
attempts were made to obtain quantitative results, in part, because of uncertainties regarding the separation efficiency of
mineral precipitates during the sonication step.
A summary of the XRD results is reported in Table 3.7. Magnetite (Fe3O4) was observed in every iron core sample.
Quartz (SiO2) was sometimes detected, particularly in samples proximal to the edges of the reactive media. Aquifer
grains had either been transported into the reactive media at some point during or after construction of the PRB, or the
core sampling may have resulted in some limited mixing of aquifer material with the reactive iron zone. Aragonite
(CaCO3) and calcite (CaCO3) were detected as minor components in some of the iron core samples. In addition to
magnetite, other iron minerals detected were green rust (GR1) and iron carbonate hydroxide. Neither siderite (FeCO3)
nor iron sulfides were detected by X-ray diffraction in any of the samples analyzed. However, using micro-analytical
methods, Furukawa et al. (2002) detected quantities of mackinawite, greigite, and pyrite in sample EC060200-1-3, as
well as ferrihydrite (Fe{OH)3) and lepidocrocite (FeOOH). Bulk X-ray analysis apparently captures the dominant mineral
phases present, but the technique was not sufficiently sensitive to identify the full range of materials that form in PRBs,
particularly those phases that have a poor degree of crystallinity. The principal authigenic phases present in the
Elizabeth City reactive media are iron carbonate hydroxide, mackinawite, and magnetite; minor mineral products are
aragonite, calcite, lepidocrocite, and green rust (GR1).
Iron carbonate hydroxide was identified and described by Erdos and Altorfer (1976) as an iron corrosion product formed
in carbonate solutions. The stoichiometry of the material is reported to be Fe9(OH),,CO3 (powder diffraction file PDF 33-
0650). The material is related to siderite, FeCO3, and the carbonate form of green rust, Fe6(OH)12CO3-x H2O. Like
siderite, iron is present only in the ferrous state in iron carbonate hydroxide, whereas in green rust, iron is present in both
the ferrous and ferric states (McGill et al., 1976). Iron hydroxide carbonate may be a precursor to carbonate green rust
and magnetite, these phases forming by partial oxidation. Iron carbonate hydroxide was also found to be a major
corrosion product at the Denver Federal Center (see Section 4.2.3) and was detected at the Moffett Field PRB (Gavaskar
et al., 2002) and the Oak Ridge Y-12 site (Liang et al., 2003).
Mixed valence iron minerals include magnetite and carbonate green-rust compounds. Green-rust compounds are iron
corrosion products that are expected to form under more reducing conditions than do ferric oxyhydroxides. Green rust
precipitation is favored under moderately alkaline conditions, and transformation of these compounds to magnetite is
expected based upon experimental evidence and thermodynamic calculations (e.g., Bonin et al., 2000). Green-rusts are
highly susceptible to oxidation in the presence of dissolved oxygen or air. Green rust structural units consist of
alternating positively charged tri-octahedral metal hydroxide sheets and negatively charged interlayers of anions (Taylor,
1973). Two types of GR are distinguishable based upon X-ray analyses: GR1 in which the distance between hydroxide
sheets is between about 0.75 and 0.80 nm (e.g., GRCO32~) and GR2 in which the distance between sheets is about
1.1 nm (e.g., GRSO42~). X-ray diffraction results are consistent with the presence of GR1 in core materials from Elizabeth
City.
39
-------
3000-
2500-
2000-
"(75 1500-
1000-
500-
a)
mid-barrier
upgradient edge
mid-barrier
upgradient edge
V-AJ
EC060300-4-5
EC060300-4-2
EC060300-5-3
EC060200-1-3
0 10 20 30 40 50 60 70 80 90
degrees 2-theta
2500-
2000-
1500-
1000-
500-
0-
upgradient edge
EC050801-3-7
EC050801-3-5
EC050801-3-3
EC050801-3-2
EC050801-3-1
0 10 20 30 40 50 60 70 80 90
degrees 2-theta
Figure 3.23 Powder X-ray diffraction data from fine-grained materials removed via sonication from cores collected at
the Elizabeth City PRB: a) core EC060300-4; b) core EC050801 -3.
40
-------
Table 3.7. Results of Powder X-ray Diffraction Analysis of Core Materials from the Elizabeth City PRB
Sample
EC060200-1-3
EC060300-5-3
EC060300-4-2
EC060300-4-5
EC050801-3-1
EC050801-3-2
EC050801-3-3
EC050801-3-7
EC050801-3-7
Major Component
Magnetite, Iron
carbonate hydroxide
Magnetite, Iron
carbonate hydroxide
Magnetite, Iron
carbonate hydroxide
Magnetite, Iron
carbonate hydroxide
Magnetite, Iron
carbonate hydroxide
Magnetite, Iron
carbonate hydroxide
Magnetite, Iron
carbonate hydroxide
Magnetite, Iron
carbonate hydroxide
Magnetite, Iron
carbonate hydroxide
Minor Component
Quartz, Calcite
Quartz, Calcite
Quartz, Aragonite,
Calcite
Quartz, Calcite
Quartz, Aragonite,
Calcite, Green rust 1
Quartz, Calcite
Calcite
Calcite
Quartz, Calcite
Trace Component
Lepidocrocite,
Mackinawite
Carbon
3.2.5
Scanning electron microscopy (SEM) was used to evaluate the morphology and spatial relationships among mineral
precipitates on the surfaces of zero-valent iron particles. In addition, energy dispersive X-ray spectroscopy (EDX) was
conducted to determine on a semi-quantitative basis the composition of surface precipitates. Samples for SEM and EDX
analysis were stored in an anaerobic glove box and then embedded in an epoxy resin. The sample mounts {1" diameter
round mounts) were ground and polished using diamond abrasives and coated with a thin layer of carbon prior to being
placed within the SEM sample chamber.
SEM photomicrographs are presented in Figure 3.24a-c for three samples from core EC060300-4 from the Elizabeth City
site collected in June 2000. These three samples were retrieved from near the upgradient aquifer/iron interface region
(EC060300-4-1, horizontal penetration ~ 2 cm; EC060300-4-3, horizontal penetration ~8 cm), and the downgradient
edge of the reactive iron media (EC060300-4-7, horizontal penetration ~40 cm). The micrographs in Figure 3.24 are
representative of particles contained in each of the samples and capture a range of magnifications from about 50x to
SOOOx.
Sample EC060300-4-1 was collected from a region near the upgradient interface and shows a fairly consistent
accumulation of mineral precipitates on the surfaces of iron particles (Figure 3.24a). In this sample, the thickness of
surface coatings on iron grains ranges from about 20 jim to 100 u.m. This surface precipitate buildup occurred over the
first four years of operation of the PRB. An average linear rate of precipitate accumulation of about 5 to 25 p,m per year
is indicated. The SEM images exhibit two types of precipitate morphology: platy/acicular aggregates and poorly
crystalline clusters (Furukawa et al., 2002). Individual particles in the platy aggregates are on the order of 10 u,m in
length. Platy or acicular textures are oftentimes the result of rapid precipitation from highly oversaturated solutions.
41
-------
a)
'1 ";,•„
.-.
| 25 kV 2000x
Figure 3.24 (continued) Scanning electron micrographs of samples from the Elizabeth City PRB: a) sample
EC060300-4-1 located near the upgradient iron/aquifer interface; b) sample EC060300-4-3 located in
the midbarrier; and, c) EC060300-4-7 located near the downgradient iron/aquifer interface.
42
-------
25 kV 350x
Figure 3.24 (continued) Scanning electron micrographs of samples from the Elizabeth City PRB: a) sample
EC060300-4-1 located near the upgradient iron/aquifer interface; b) sample EC060300-4-3 located in
the midbarrier; and, c) EC060300-4-7 located near the downgradient iron/aquifer interface.
43
-------
c)
15kV 75x 150nm
15kV 200x SOiim
, *•:-' v.v: s, v
. '•• t. -,T »»a?.
500x
Figure 3.24 (continued) Scanning electron micrographs of samples from the Elizabeth City PRB: a) sample
EC060300-4-1 located near the upgradient iron/aquifer interface; b) sample EC060300-4-3 located in
the midbarrier; and, c) EC060300-4-7 located near the downgradient iron/aquifer interface.
44
-------
Such conditions might be expected in the region where ground water passes from the aquifer into the reactive zone, i.e.,
where steep geochemical gradients are present (for example in pH and Eh). These observed morphologies are
consistent with previous studies (e.g., Roh et al., 2000; Phillips et al., 2000). These previous studies found that acicular/
platy aggregates in zero-valent iron PRBs were largely composed of green rust minerals, goethite, lepidocrocite, and
calcium carbonates. The poorly crystalline clusters more typically were composed of mackinawite and poorly
crystallized iron oxyhydroxides (Phillips et al., 2000). Poorly crystalline clusters are typically found immediately on the
zero-valent iron surfaces, with the elongated particles apparently growing on the clusters. In other cases, however, the
poorly crystalline clusters are found together with the elongated particles and not directly associated with the iron
surfaces.
Sample EC060300-4-3 was collected near upgradient aquifer/PRB region but at a penetration depth further into the PRB
compared to sample EC060300-4-1. Sample EC060300-4-3 shows the same particle morphologies as were noted in
sample EC060300-4-1. However, the regularity of surface precipitates is less pronounced and the thickness of the
surface precipitate layer is thinner, ranging from 0 to about 50 (im. In this sample, an average linear rate of precipitate
accumulation of about 0 to 13 u,m per year is indicated (Figure 3.24b). In sample EC060300-4-7, collected near the
downgradient aquifer/iron interface, the surface precipitate thickness is very thin if present at all (Figure 3.24c). In most
cases a 1- to 4-jim layer was observed on the iron grains in this sample collected near the downgradient edge of the
reactive media. This observation suggests that after four years of ground-water exposure, iron particles free from
precipitate coverage are still present at the Elizabeth City site. The PRB should still remain reactive, which is consistent
with the previously described trends in geochemical parameters and contaminant concentrations.
Energy dispersive analyses of iron grains from the Elizabeth City PRB indicate the presence of iron (~97 wt%), silicon
(~2.5%), and occasionally Mn (<0.7 wt%) and Cr (<0.6 wt%). Peerless iron is also known to contain concentrations of
carbon and other trace elements, including S, P, Ni, V, Mo, Ti, and Cu. Because the samples were coated with carbon,
this element could not be semi-quantitatively determined; other trace elements are present at concentrations below the
operational detection limit of the EDX method. Oxygen was not detected on freshly polished surfaces of zero-valent iron.
However, the coatings of surface precipitates are enriched in oxygen and comparatively depleted in iron relative to the
composition of fresh zero-valent iron (Figure 3.25). Typically the most oxygen-enriched regions are those farthest away
from the zero-valent iron surface. Particles with a platy morphology were found to contain iron (68.9 ± 8.2 wt%, n=25),
oxygen (28.3 ± 7.5 wt%, n=25), silicon (2.4 ± 4.2 wt%, n=25), sulfur (0.2 ± 0.6 wt%, n=25), and manganese
(0.5 ± 0.5 wt%, n=25) (Figure 3.26). Chromium was never detected in spot analyses of the platy particles, and calcium
was detected in only 6 of 25 analyses, always at levels below 1 wt%. The poorly crystalline clusters are similar to the
platy particles with respect to iron concentrations (64.5 ± 8.0 wt%, n=30) and are broadly comparable in terms of the
element distributions of Si, S, Mn, and Ca. Chromium was detected in 10 of 30 of the spot analyses of the poorly
crystalline clusters in concentrations up to about 2 wt%.
Chromium was detected by EDX spot or small area analysis in 72 separate measurements. Concentrations of chromium
ranged from 0.1 wt% to 1.8 wt%. The micro-scale measurements are much greater than bulk-scale Cr concentrations
determined by acid leaching large (~1-5 gm) quantities of material. Pearson correlation coefficients presented in
Table 3.8 indicate significant correlation of chromium abundance with S and Mn. The high degree of correlation with Mn
suggests that Cr uptake is in some way tied to Mn behavior; either they are present together in some discrete phase or
as a co-precipitate. This observation is somewhat surprising as most laboratory studies have concluded that Cr is
removed from solution through the formation of a solid solution or by adsorption of Cr(lll) onto iron oxyhydroxide surfaces
(Powell et al., 1995; Blowes et al., 1997; Pratt et al., 1997).
XPS scans show that iron particle surfaces from Elizabeth City contain C, O, Fe, Si, S, Mg, Ca, Mn, and N. The XPS
data indicate a surface layer dominated by iron oxyhydroxides, an intermediate layer of iron oxide, and finally, zero-
valent iron at the greatest sputtering depths. Surface carbon is present predominately as carbonate with some detected
hydrocarbon (binding energy 284.6 eV). The oxidation state of sulfur is predominantly present as sulfide (-2) but with
minor amounts of sulfate (+6). Surface enrichment in the elements Ca, Mg, S, and Si are consistent with observed
decreases in ground-water concentrations of these elements. Chromium was sometimes detected by XPS in iron
samples from Elizabeth City.
3.2.8
From 1999 to 2001,117 samples were collected from the Elizabeth City site for phospholipid fatty acid (PLFA) extract
characterization. Samples were collected from regions within the reactive iron media and from adjacent regions of the
upgradient and downgradient aquifer. Samples for PLFA analysis were frozen immediately after collection from the
subsurface and shipped frozen to Microbial Insights (Rockford, Tennessee). The complete PLFA data set from the
Elizabeth City site is shown in Table B1 (Appendix B) and summarized in Table 3.9.
Biomass contents spanned several orders of magnitude from <1 to 2614 picomoles per gram (dry weight basis), or from
about 2.5x103 to 5.23x107 cells per gram of dry material. The biomass is dominated by Prokaryote PLFA. The highest
biomass concentrations were found near the upgradient aquifer/iron interface region, in the same region of the reactive
45
-------
100-
95-
90-
85-
80-
75-
70-
65-
60-
55-
50-
45-
40
• Iron particles
• Precipitates directly on Iron particles
• Platy particles
Amorphous particles
10 15
20 25 30
O, Wt%
35 40 45 50
Figure 3.25 Iron concentration versus oxygen concentration in iron grains and surface precipitates (SEM-EDX); Fe-
O compositions noted for wustite (W, FeO), magnetite (M, Fe O ), hematite (H, Fe O ), goethite (G,
FeOOH), and ferrihydrite (F, Fe(OH)3).
Table 3.8. Pearson's Correlation Matrix of Element Concentrations Determined by SEM-EDX Analysis
O
Si
s
Cl
Mn
Fe
Cr
Ca
O
1
0.712
0.031
0.242
-0.127
-0.747
0.057
0.597
Si
1
-0.126
0.281
-0.082
-0.532
0.014
0.471
S
1
0.177
0.346
-0.658
0.254
-0.119
Cl
1
0.458
-0.392
0.292
0.029
Mn Fe
1
-0.237 1
0.655 -0.269
-0.187 -0.351
Cr Ca
1
-0.162 1
media where enrichments in inorganic precipitates are observed (Figure 3.27 and 3.28). Downgradient regions tend to
be comparatively depleted in microbial biomass (Figure 3.28). The lower counts associated with the mid-barrier and
downgradient samples suggest that the environment at these locations is more challenging to bacterial growth and
survival. Examining the geochemical conditions associated with these locations supports this hypothesis. Figures 3.5
and 3.15, for example, indicate a decrease in biologically available electron acceptors such as cis-DCE and sulfate in
mid-wall and downgradient locations. The higher pH and lower availability of electron acceptors would also tend to
create a more severe environment for bacterial growth.
The highest biomass contents detected were found in samples collected in 2001, four years after installation of the iron
wall (sample EC050801-5-2). However, nearly equivalent values were detected in samples collected in 1999 (sample
EC90903), only two years after installation. These data suggest non-constant microbial growth rates, or they may
46
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3.9. Summary of PLFA Data from the Elizabeth City PRB
Number of samples
Average PLFA concentration
(pmoles/g)
PLFA range
PLFA Structural Groups
average % (range %)
Monoenoic
Found in Gram-negative
bacteria
Terminally Branched
Saturated
Found in many Gram-positive
bacteria, and in some Gram-
negative bacteria
Branched Monoenoic
Common in obligate
anaerobes, such as sulfate-
reducing and iron-reducing
bacteria
Mid-Chain Branched
Saturated
Common in actinomycetes and
sulfate-reducing bacteria
Normal Saturated
Ubiquitous in prokaryotes and
eukaryotes
Polyenoic
Found in fungi, protozoa,
algae, higher plants, and
animals
Fe(0)
Upgradient
Elizabeth City
27
611
6-2614
36.9
(14.0-73.5)
19.1
(5.3-42.1)
11.5
(<1-30.3)
7.8
(<1-38.9)
22.6
(5.2-74.5)
2.0
(<1-49.6)
Fe(0)
Downgradient
Elizabeth City
33
26
<1-260
35.9
(<1-74.8)
4.1
(<1-28.3)
3.1
(<1-37.8)
0.6
(<1-7.9)
50.4
(<1-100)
2.9
(<1-53.7)
Aquifer
Elizabeth City
46
33
<1-348
28.9
(<1-63.0)
4.5
(<1-19.2)
3.9
(<1-29.8)
2.4
(<1-17.1)
59.1
(15.6-100)
1.3
(<1-9.5)
suggest that comparisons of microbial biomass concentrations must be made in the context of the sample location
relative to the ground-water plume (i.e., the input of electron acceptors). The enrichments in microbial biomass at 5 to
6 meters below ground surface correspond to the depth where the highest concentrations of terminal electron accepting
species enter the reactive media via ground-water transport (Figure 3.27).
PLFA profiles from the Elizabeth City site are enriched in fatty acid biomarkers indicative of anaerobic sulfate- or iron-
reducing bacteria (Dowling et al., 1986; Edlund et al., 1986; Tunlid and White, 1991; Parkes et al., 1993). The high
proportions of terminally branched and branched monoenoic PLFA specifically indicate anaerobic metabolism. Termi-
nally branched PLFA are typical of Gram-positive bacteria, but can also be present in the cell membranes of some
anaerobic Gram-negative bacteria. However, because high proportions are present of branched monoenoic PLFA
47
-------
10-
2-
0-
Iron
Surfaces
Platy
Particles
Amorphous
Particles
Figure 3.26 Element concentrations in surface precipitates from the Elizabeth City PRB; iron surfaces=precipitates
directly in contact with iron grains.
indicative of anaerobic metal reducing bacteria, the terminally branched PLFA are likely to be mainly from sulfate or iron
reducing bacteria. Where biomass is most concentrated (i.e., near the upgradient aquifer/iron interface), the distribution
of PLFA overall appears to be distinct from the PLFA distribution observed in the native aquifer materials (Figure 3.29).
Near the upgradient aquifer/iron interface, the proportion is greater of branched monoenoic PLFA and PLFA indicative of
sulfate-reducing bacteria compared to the PLFA signature of native aquifer materials (Figure 3.29). Where biomass is
least concentrated (i.e., near the downgradient aquifer/iron interface), the distribution of PLFA overall appears to match
the PLFA distribution observed in the native aquifer materials (Figure 3.29).
3.3 Summary of Results from the Elizabeth City Site
Results of the long-term performance evaluation at the Elizabeth City site indicate that the reactive barrier there
continues to remove contaminants from ground water after five years of operation. The PRB at Elizabeth City is
expected to retain an effective level of reactivity and hydraulic performance for at least another five year time period. The
salient results of the Elizabeth City site study are summarized below:
• Removal of contaminants, Cr, TCE, cis-DCE, and VC, continues after five years of PRB operation. In all
cases, chromium concentrations have been reduced to below the MCL, and in the majority of sampling
events, Cr was undetected in monitoring wells located downgradient from the PRB. Concentrations of
volatile organic compounds have been significantly reduced, but TCE concentrations above the MCL have
been observed in some downgradient wells.
• After five years, ground water in the PRB is moderately alkaline (pH>9) and moderately reducing (Eh<-100
mV). Time trends in pH suggest quasi-steady-state conditions. Time trends in Eh, however, suggest that
the PRB is gradually losing the capacity to produce reducing conditions due to progressive exposure to
ground water. Time trends in specific conductance values indicate that influent solutes continue to be
48
-------
PLFA
8
x;2614 pmoles/g
-0.6 -0.2 0 0.2
Position, m
<50 pmoles/g
50-500 pmoles/g
500-1000 pmoles/g
1000-2000 pmoles/g
>2000 pmoles/g
Figure 3.27 Cross-sectional profile showing concentration distribution of biomass (from PLFA data) in picomoles per
gram, Elizabeth City PRB (June 2002).
49
-------
CD
CT
CD
16-
14-
12-
10-
6-
4-
2-
0-
20-
16-
12-
8-
4-
I Aquifer)
CD
CT
CD
Fe(0) downgradient zone
1 H
Fe(0) upgradient zone
1000 2000
PLFA, picomoles/g
3000
Figure 3.28 Histograms of microbial biomass concentrations (from Elizabeth City PLFA data) in picomoles per gram
in aquifer materials, iron from near the downgradient aquifer/iron interface, and iron from near the
upgradient aquifer/iron interface.
50
-------
Aquifer
60%
29%
Fe(0)
Downgradient
0.6%
3.1%
1.3%
4.5% ^H Gram+/Gram- (Ter. Br. Sat.)
i i Gram- (Monoenoic)
I | Anaerobic (Br. Mono.)
SRB, Actinomycetes
49%
] Genera Nsats
] Polyenoic
^^^^ 2.5%
36% 4.1%
Fe(0)
Upgradient
12% 7.8%
19%
Figure 3.29 Pie graphs showing structural distribution of PLFA compounds (average values) at the Elizabeth City
site.
51
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removed within the reactive medium and partitioned into the solid phase after five years of PRB operation.
The geochemistry of ground water downgradient of the PRB is impacted by the PRB, in that higher pH and
lower Eh conditions are observed in wells located within 2 m of the downgradient iron/aquifer interface.
The Elizabeth City PRB has consistently removed inorganic carbon, sulfate, calcium, magnesium, silicon,
and nitrogen from influent ground water. These components have either been precipitated out in the PRB,
adsorbed to iron granules or secondary precipitates within the PRB, or have been chemically transformed by
biotic or abiotic processes.
Mineralogical characterization of soil core materials indicates the formation of calcite/aragonite, iron
carbonate hydroxide, magnetite, lepidocrocite, mackinawite, and carbonate green rust in the PRB at
Elizabeth City. Inorganic carbon is present in several calcium- and iron-containing minerals. Sulfur is
dominantly present in the iron sulfide mackinawite. Mineral precipitation mainly occurs near the upgradient
edge of the PRB, although there is an indication that a precipitation front is progressively moving through the
PRB. Iron core collected near the downgradient edge of the PRB contains very little if any mineral
precipitate mass. After five years of operation, than 10% of the available pore space has been lost due
to mineralization near the upgradient edge. Near the downgradient edge, <1 % of the available pore space
has been lost.
Microscopic characterization of core materials indicates that mineral accumulation is occurring mainly on the
surfaces of iron granules. After five years, coverage of iron granules near the upgradient edge is regular and
approximately 20 to 100 urn thick. Near the downgradient edge, coverage of the iron grains is
consistent and where present mineral coatings are generally <5 urn thick. Although the available reactive
surface of Fe° has been reduced through time, some of the secondary mineral precipitates identified
(magnetite, green rust, mackinawite) also support contaminant transformation and uptake, thus potentially
compensating for the loss in iron metal reactivity due to surface precipitation.
Microbial characterization results, based on PLFA profiles, from the Elizabeth City PRB and adjacent aquifer
materials showed a diverse microbiological community dominated by Gram-negative bacteria. Iron core
samples from near the upgradient edge of the PRB are typically enriched in microbial biomass (up to
5.23x107cells/g) and contained elevated proportions of biomarkers indicative of metal-reducing and sulfate-
reducing bacteria. Aquifer materials (up to 6.96x106 cells/g) and iron from near the downgradient PRB edge
(5.20x106 cells/g) were comparatively depleted in total biomass and in biomarkers indicative of metal-
reducing and sulfate-reducing bacteria.
52
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4.0
4.1
Ground-water monitoring wells in and around gates 1, 2, and 3 at the Denver Federal Center (DFC) were sampled on an
annual basis starting in May of 1999, approximately 2.5 years after the f unnel-and-gate system was installed. Up to 20
monitoring wells were sampled to obtain contaminant and geochemical profiles in upgradient, iron media, and
downgradient positions. The Federal Highway Administration provided a more extensive ground-water data set,
covering the entire history of the DFC PRB (FHWA, 2002 pers. common.; Pacific Western Technologies, 2000). These
secondary data were collected under an approved QA/QC program (USGS, 1999). The locations of the monitoring wells
in relation to the reactive iron media in 1-3 at the DFC are shown in Figures 4.1-4.3.
Methods of ground-water and soil core sampling, preservation and analysis, in addition to Quality Assurance/Quality
Control procedures used in this five-year investigation of PRB performance, are described and discussed in Volume 2 of
this EPA Report series.
4.1.1 I
Time-dependent concentrations of volatile organic compounds (1,1-DCE, 1,1,1-TCA, TCE, and 1,1-DCA) in monitoring
wells located in upgradient, iron wall, and downgradient positions near gate 1 are shown in Figure 4.4. Data plotted in
Figures 4.4, 4.5, and 4.7 were provided by FHWA. It should be noted that not all monitoring data collected at the DFC
are discussed in this report. The data presented correspond to wells that overlap between EPA's performance study and
monitoring efforts conducted by FHWA and GSA. The concentrations of 1,1-DCE and 1,1,1-TCA have gradually
decreased with time in upgradient well GSA-21 (Figures 4.4a). This trend in contaminant concentrations suggests that
natural attenuation processes are taking place in regions of the plume upgradient from the southern most extension of
the f unnel-and-gate system at the DFC. Concentrations of TCE and 1,1-DCA in ground water entering gate 1 at the DFC
have remained fairly constant since January 1998 at 23 ± 7 (ig/L and 7 ± 2 u,g/L, respectively. Concentrations of vinyl
chloride and cis-DCE have not been detected in upgradient monitoring points in the vicinity of gate 1.
Within the iron media of 1 (well C1-I2), concentrations of TCE and 1,1,1-TCA have remained at or below nominal
quantification limits (Figure 4.4b). Concentrations of 1,1-DCE were not detected until November 1999. At that time,
concentrations of 1,1 -DCE began to increase at a time-averaged rate of about 5 u,g/L per year. Similarly, concentrations
of 1,1 -DCA were not detected until November 1999. However, since that time 1,1 -DCA has remained at a fairly constant
value of 7 ± 0.8 fig/L Concentrations of vinyl chloride and cis-DCE have not been detected in monitoring wells located
within the reactive media of gate 1. Although lower contaminant concentrations are present, parallel trends in
contaminant levels are observed in well C1-I2, located approximately 3 m to the south of well C1-I1 along the mid-point
axis of the reactive media in gate 1 (see Figure 4.1).
Based on laboratory studies to evaluate contaminant removal rates in the presence of zero-valent iron, degradation rates
of chlorinated ethenes and ethanes decrease in the order 1,1,1-TCA>TCE>1,1-DCE>cis-DCE>VC (see Johnson et al.,
1996). In general, with each successive dehalogenation the degradation reaction proceeds more slowly (Matheson and
Tratnyek, 1994; but also see Arnold and Roberts, 2000). In a performance scenario where zero-valent iron progressively
loses reactivity over time relative to an initial reactivity defined by laboratory batch or column studies, it might be
expected that less reactive contaminants such as VC, cis-DCE, and 1,1 -DCE would appear or "break through" prior to
more rapidly degraded compounds such as TCE and 1,1,1-TCA. These trends are in fact observed in gate 1 at the DFC.
It should be noted that monitoring points C1-I1 and C1-I2are located at the approximate mid-point of the reactive media
so that contaminants should continue to degrade in the downgradient direction.
Ground-water seepage velocities in gate 1 have been measured using heat-flow sensors and range from about 0.40 to
0.82 m/d (McMahon et al., 1999; Pacific Western Technologies, 2000). Based on this range of flow rates, an average
saturated thickness of 4.5 m, and a saturated gate throughput area of 55 m2, gate 1 has removed approximately
0.8-1.6 kg of TCE, 2-4 kg of 1,1,1-TCA, and 3.2-6.4 kg of 1,1-DCE over the first five-year period of operation. By
February 1998, a clean front was observed in downgradient well GSA-20. At that time, concentrations of 1,1-DCE and
1,1, DCA increased to above detectable limits and have remained at detectable levels to the present time. The reasons
for the increase in the concentrations of these compounds is not certain but would appear to relate to a loss of reactivity
or change in residence time of contaminants in the reactive media.
53
-------
Cl-3-71701
o
Groundwater flow
e GSA-21
Cl-2-71701
O Oci-1-71701
O
Cl-2-71000
C1-GU1
1.5m
© Monitoring well location
O Angle or vertical
coring position
GSA-20
C1-3-71100Q
C1-GD1C1-4-71801
Figure 4.1. Coring locations and monitoring well locations at the Denver Federal Center, gate 1 (plan view).
54
-------
Groundwaterflow
GSA-26
USGS-10
USGS-12
C2-GU2
C2-12-71300
C2-13-71300
C2-14-71300
C2-1-71901 "
C2-2-71901
C2-3-71901
1.5 m
© Monitoring well location
O Angle or vertical
coring position
_
o
TO
0
O.
CO p
(D
-------
Groundwaterflow
GSA-31
8 C3-GU2
© Monitoring well location
O Angle or vertical
coring position
C3-2-71801
O
GSA-30
©
Figure 4.3. Coring locations and monitoring well locations at the Denver Federal Center, gate 3 (plan view).
56
-------
c
O
c
CD
O
8
180.
160-
140-
120-
100-
80-
60-
40-
20-
0
.^i-rTTrrT
-.T-T-T-l
GSA-21
Gate 1
Upgradient
10/10/1995 2/21 /1997 7/6/1998 11 /18/1999 4/1 /2001 8/14/2002
40
c
g
15
i
c
CD
O
C
O
O
c
O
c
CD
O
8
30-
20-
10-
0-
b)
*•«»*»*»*»»*»'->
C1-I1
Gatel
Fe(0)
10/10/1995 2/21/1997 7/6/1998 11/18/1999 4/1/2001 8/14/2002
20
16-
12-
4-
c)
_: u '_j '_' -^
T T-T-T T-T T'T'T
GSA-20
Gate 1
Downgradient
10/10/19952/21/19977/6/1998 11/18/19994/1/2001 8/14/2002
Figure 4.4 Concentrations of contaminants through time in monitoring wells from the Denver Federal Center, gate 1
(data from FHWA): a) well GSA-21 (upgradient); b) well C1-I1 (iron wall); c) well GSA-20
(downgradient). Symbol key shown in Figure 4.4a applies to each graph; note change in concentration
plotting range to show trends through time.
57
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4.1.2 2
Gate 2 is located approximately 120 meters north of gate 1. Of the four treatment cells at the DFC, the highest
concentrations of total volatile organic compounds are found in the vicinity nearby gate 2. Since November 1996,
concentrations of 1,1 -DCE and 1,1,1-TCA have gradually decreased in the upgradient well, GSA-26, to values of about
110 jig/L and 50 |ig/L, respectively (Figure 4.5a). In this same well, concentrations of TCE, 1,1-DCA, and cis-1,2-DCE
have been steady since November 1996 at 72 ± 12 fig/L, 8 ± 1 u.g/L, and 1 ± 0,5 u.g/L, respectively (Figure 4.5a). The
highest TCE concentrations observed, in either upgradient or downgradient monitoring locations, are in downgradient
well GSA-25 (Figure 4.5c). A maximum TCE concentration of 600 jig/L was measured in ground water collected from
GSA-25 in January 1997. From January 1997 to November 1998, TCE concentrations rapidly dropped in GSA-25, and
since November 1998, TCE and all other volatile organic compounds have remained at relatively constant concentra-
tions (Figure 4.5c).
Because TCE concentrations are comparatively low in wells located within the iron media of gate 2 and in the gravel
immediately downgradient of gate 2, the TCE observed in downgradient well GSA-25 is thought to be derived from
residual bedrock contamination from a proximal release of TCE (Pacific Western Technologies, 2000). Yet, concentra-
tions of 1,1 -DCE have been consistently present in wells both within iron media of gate 2 and in downgradient pea gravel
wells (Figure 4.5b). The increase beginning in November 1998 of 1,1-DCE levels both in downgradient well GSA-25 and
in the iron well C2-I2 corresponds to the installation of a distribution trench placed on the downgradient side of gate 2.
The trench was installed with the goals of increasing the flow through gate 2 and lowering the hydraulic head differential
across the gate. Although the timing of trench construction matches the point at which 1,1-DCE concentrations increase,
the mechanism linking these events is uncertain. Alternatively, the increasing 1,1-DCE levels could be from incomplete
degradation within the iron media due either to loss of iron reactivity or to changing flow regimes within gate 2 (decrease
in residence time). Iron reactivity tests and mineral precipitate/microbial biomass characterization studies on the iron
media in gate 2 are discussed in following sections of this report. Equally equivocal is the trend of decreasing
concentrations in C2-I2 beginning in September 2001 (Figure 4.5b), which would be an unexpected outcome if
channeling was occurring or if the iron media had lost reactivity.
A previous study proposed that ground-water mounding upgradient of gate 2 was likely driving flow underneath of the
funnel-and-gate system near gate 2 (McMahon et al., 1999). Multi-level sampling using low-volume diffusion samplers
in well GSA-25 shows that TCE concentrations in this well are highly depth dependent with the greatest concentrations
detected near the bottom of the screened interval (Figure 4.6). These concentration trends cannot be used to reliably
confirm or refute underflow of contaminants beneath the gate. The necessary hydraulic data are not available to make
this assessment. However, examination of [TCE]/[1,1-DCE] ratios suggests that ground water upgradient of gate 2
(1,1-DCE-rich) is distinct from and not likely to be a source of ground water present in downgradient well GSA-25.
4.1.3 3
The concentrations of volatile organic compounds in monitoring wells around gate 3 of the DFC are shown in Figure 4.7.
In general, the concentrations of individual contaminants are at levels below 5 p,g/L in well C3-12 located within the
treatment cell, with the exceptions of late-summer sampling events in August 2000, September 2001, and August 2002.
In gate 3, temporal trends in contaminant distributions are clearly intermittent. Concentrations of cis-DCE, 1,1 -DCE, VC,
and 1,1-DCA in upgradient well GSA-31, for example, spiked in August of 2000. At the same time, elevated contaminant
concentrations are found in well C3-I2 located within the reactive media. However, co-temporal spikes are not observed
in downgradient well GSA-30. As in gate 2, contaminants detected in downgradient monitoring points are thought to be
derived from residual bedrock contamination (Pacific Western Technologies, 2000). Correlations have been noted
between high concentrations of volatile organic compounds and low ground water levels, but the mechanism linking
these observations is unclear.
4.1.4
At the Denver Federal Center, trends in pH have followed consistent patterns from May 1999 to July 2001. Gates 1
through 3 show comparable trends in pH from upgradient to downgradient sampling locations (Figure 4.8). Trends in
contaminant degradation behavior among the reactive cells do not clearly correlate with pH, i.e., pH values in gate 2 are
similar to pH values measured in gate 1 and gate 3. One notable feature is the continued decrease in pH of about 0.1 pH
units per year in ground water from the iron media in gates 1, 2, and 3. This trend is marginally significant with respect
to the precision and accuracy of the pH measurements but might be related to an overall decrease in residence time of
ground water in the reactive media, related to pore-infilling by mineral precipitates and microbial biomass.
The specific conductance (SC) of ground water at the Denver Federal Center shows variable patterns (Figure 4.9). In
gate 1, SC values in the reactive iron media decrease by an average of about 31% relative to upgradient ground water;
in gate 3 SC values in the reactive media decrease by about 28% relative to upgradient ground water. As described
previously, decreases in SC are expected as ground water passes through zero-valent iron reaction zones due to
mineral precipitation. On the other hand, trends in SC in gate 2 are anomalous. In 1999 and 2000, SC values in the iron
58
-------
O
"CD
8
Ł=
8
250.
200-
150-
100-
50-
0.
I ,,
GSA-26
Gate 2
Upgradient
10/10/1995 2/21/1997 7/6/1998 11/18/1999 4/1/2001 8/14/2002
150.
120-
90-
8 60-|
O
O
30-
0-
b)
C2-I2
Gate 2
Fe(0)
10/10/1995 2/10/19976/10/1998 10/10/19992/10/2001 6/10/2002
§]
600-
500-
400-
= 300-
Ł=
CD
O
c
O
O
200-
100-
0-
GSA-25
Gate 2
Downgradient
10/10/1995 2/21/19977/6/1998 11/18/19994/1/2001 8/14/2002
Figure 4.5 Concentrations of contaminants through time in monitoring wells from the Denver Federal Center, gate 2
(data from FHWA): a) well GSA-26 (upgradient); b) well C2-I2 (iron wall); c) well GSA-25
(downgradient). Symbol key shown in Figure 4.5c applies to each graph; note change in concentration
plotting range to show trends through time.
59
-------
a)
15-
f 20-
_cŁ
Q.
0
~° 25-
30-
(
10-,
b)
15-
O)
'o.
0
~° 25-
30-
—•— 1,1-DCE
—•—1,1,1-TCA
TCE
GSA-26 (upgradient, gate 2)
N L
»
/ .X
< f
) 20 40 60 80 100 120 140 160 180
concentration, u,g/L
— • — sulfate
• calcium
iron
GSA-26 (upgradient, gate 2)
i
/ i
\m \f
15-
20-
25-
30-
(
10-,
15-
20-
25-
1
30 -I
—•—1,1-DCE
—•—1,1,1-TCA
TCE
GSA-25 (downgradient, gate 2)
r /
• {
u
J 20 40 60 80 100 120 140 160 180
concentration, u,g/L
— •— sulfate
— •— calcium
iron
GSA-25 (downgradient, gate 2)
\ \
\ !
\ I
0 50 100 150 200 250 300 350 400 450 0 50 100 150 200 250 300 350 400 450
concentration, mg/L concentration, mg/L
Figure 4.6 Depth-resolved concentrations of a) contaminants (|^g/L) and b) sulfate, calcium, and iron (mg/L) in wells
GSA-26 and GSA-25 from the Denver Federal Center (gate 2). Data collected in May 1999.
60
-------
70-
65-
60-
55-
50-
45-
=J 40-
S 35:
Ł 30-
1 25-
^ 20-
03 H r-
o 15-
c
o 10-
5-
0-
-5-
3) _«_1;1.DCE
— •— 1,1,1-TCA
TCE
-T-1.1-DCA
cis-1,2-DCE
A
-A\"~B" ra
T T"''T^i i \
V -^V j^fej
• •-•-• »-?5^'j *^u*-»o IV**^*— *
GSA-31
GateS
Upgradient
10/10/1995 2/21/1997 7/6/1998 11/18/1999 4/1/2001 8/14/2002
30
25-
I
"& 20-
c"
g
i 15"
"c
0)
g 10-
o
o
5-
0-
6/1 6/1 Ł
40-
35-
=J 30-
0)
d 25-
g
'S 20-
"c
S 15-
c
o
0 10-
5-
0-
-5-
b)
1
ft I
ii II
1 1 1
1 1 I4
II H
1 1 If T
u^Jw
^;^^^^,.H,,^...^_ .,,„..
96 10/29/1997 3/13/1999 7/25/2000 12/7/2001
c)
/:::
«HM»H^I*a«»H*P»fŁ»««
C3-I2
GateS
Fe(0)
GSA-30
GateS
Downgradient
10/10/1995 2/21/1997 7/6/1998 11/18/1999 4/1 /2001 8/14/2002
Figure 4.7 Concentrations of contaminants through time in monitoring wells from the Denver Federal Center, gate 3
(data from FHWA): a) well GSA-31 (upgradient); b) well C3-I2 (iron wall); c) well GSA-30
(downgradient). Symbol key shown in Figure 4.7a applies to each graph; note change in concentration
plotting range to show trends through time.
61
-------
pH 2001
pH 2000
pH 1999
Upgrd.
Gate 3 Fe(0)
Dwngrd.
Upgrd.
Gate 2 Fe(0)
Dwngrd.
Upgrd.
Gate 1 Fe(0)
Dwngrd.
6
PH
10
Figure 4.8 Average pH values through time in wells from upgradient, iron wall, and downgradient positions relative
to gate 1, gate 2, and gate 3 at the Denver Federal Center.
media were actually greater than the values obtained from upgradient ground water. In 2001, the trend reversed to what
is considered to be normal behavior, i.e., SC values were low within the reactive media (Figure 4.9). The high values of
SC in Cell 2 in 1999 and 2000 indicate that the iron media in gate 2 was a source of dissolved solutes rather than a sink.
Such anomalous trends in SC may be indicative of decreased performance or at least of non-typical reactive behavior.
In gate 3, downgradient ground water has a significantly higher SC than ground water upgradient and within the
treatment cell. This result suggests that the ground-water chemistry and contaminant distributions in ground-water from
well GSA-30 are not greatly influenced by water emerging from the reactive cell (gate 3).
At the Denver Federal Center, ground-water upgradient from the funnel-and-gate system is progressively more reducing
moving northward from gate 1 to gate 3. Fairly typical trends in Eh values are observed in gate 1 (Figure 4.10). Eh
values in ground water collected from upgradient and downgradient compliance wells range from about -75 mV to +150
mV. In the reactive iron media of gate 1, Eh values are negative (-175 to -250 mV), and in downgradient locations, Eh
values generally rebound to positive values. Somewhat lower and more variable Eh values are apparent in gate 2,
reinforcing the overall anomalous behavior in this iron wall (Figure 4.10).
4.1.5 Hydrogen Gas Concentrations
The concentration of dissolved hydrogen gas is another key indicator of redox conditions that may be useful in PRB
monitoring programs. Figure 4.11 shows trends in dissolved hydrogen concentrations in gate 1 at the DFC.
Comparatively high concentrations of hydrogen are observed within the reactive media of gates 1-3. High dissolved
hydrogen concentrations are an expected consequence of iron corrosion. Hydrogen concentrations are the greatest in
62
-------
Upgrd.
Gate 3 Fe(0)
Dwngrd.
Upgrd.
_ . _
uate i Fe(0)
Dwngrd.
Upgrd.
r* i
bate 1 Fe(0)
Dwngrd.
i 1 SC 2000
^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^M
1
1
1
1
|
^^^^^^^^^^^^^"
1
0 200 400 600 800 1000 1200 1400 1600 1800 2000
Specific Conductance, |aS/cm
Figure 4.9 Average specific conductance values (u,S/cm) through time in wells from upgradient, iron media, and
downgradient positions relative to gate 1, gate 2, and gate 3 at the Denver Federal Center.
I I Eh 2000
Upgrd.
Gate 3 Fe(0)
Dwngrd.
Upgrd.
Gatel Fe(0)
Dwngrd.
-0.3 -0.2 -0.1 0.0 0.1 0.2
Eh, V
0.3
Figure 4.10 Average Eh (V) values through time in wells from upgradient, iron media, and downgradient positions
relative to gate 1, gate 2, and gate 3 at the Denver Federal Center.
63
-------
-5.5-
-6.0-
-6.5-
-7.0-
_03
O
X
D)
-8.0-
o -8.5-
-9.0-
-9.5-
-10.0-
GateS
Fe(0)
Gate 1
1999
Upgradient Upgradient Fe(0) Downgradient Downgradient
peag ravel peag ravel
Position
Figure 4.11 Concentrations of dissolved hydrogen (log molar) as a function of sampling position and time in gate 1 at
the Denver Federal Center. Also shown are the concentration ranges of dissolved hydrogen measured
in the iron media in gate 2 and gate 3.
the reactive media of gate 3 and the least in gate 2. In gate 1, measured concentrations of hydrogen in the iron media
and downgradient positions have progressively decreased from 1999 to 2001. The lowest hydrogen levels were
observed in gate 2, which is consistent with the lower redox potentials indicated by platinum electrode measurements in
this iron cell. The decreased reducing potential in gate 2 indicated by Eh and H2 measurements seems to correspond
with the lower degree of contaminant removal in this reactive cell. Dissolved H2 concentrations observed within the
reactive iron media correspond to equilibrium hydrogen gas partial pressures of about 0.05 to 1 mbar. In all cases, H2
concentrations within the iron zones are greater than those typically encountered in methanogenic aquifers (5 to 30 nM,
see Chapelle et al., 1996).
4.1.6 Dissolved Cations and An ions
The cation compositions of ground water upgradient from gates 1, 2, and 3 are broadly comparable. On a molar basis,
sodium is the most abundant cation, followed by calcium, magnesium, and potassium. Cation concentrations in ground
water from upgradient wells (GSA-21, GSA-26, GSA-31) range from 5.7-7.9 mM Na, 2.7-2.9 mM Ca, 0.8-1.8 mM Mg,
and about 0.01 mM K. Ground water upgradient of gate 1 is slightly more enriched in Na, but depleted in Ca and Mg
compared to ground water from regions upgradient of gate 3 (Figures 4.12-4.14). In the reactive cells, concentrations of
calcium and magnesium are greatly reduced compared to upgradient wells, whereas concentrations of sodium are
largely unchanged relative to upgradient regions. Potassium concentrations increase slightly in the upgradient pea
gravel probably due to the dissolution of potassium-bearing aluminosilicates that are present in the pea gravel material.
From 1999 to 2001, average reductions in the concentrations of calcium and magnesium between upgradient and mid-
wall positions were greater than 95% and 75%, respectively, in gate 1. Similarly, average reductions in the
concentrations of calcium and magnesium between upgradient and mid-wall positions were greater than 95% and 50%,
respectively, in gate 2 between 1999 and 2001.
64
-------
DFC 1
UpgrcL Upgrd. PG Fe(0) Dwngrd. PG Dwngrd.
Upgrd. Upgrd. PG Fe{0) Dwngrd. PG Dwngrd.
20-,
18-
16-
14-
12-
10-
8-
6-
4-
2-
0-
1
1
T
I
I
1
f
1
Upgrd- PG Fe{0) Dwngrd- PG Dwngrd-
Upgrd- Upgrd- PG Fe{0) Dwngrd- PG Dwngrd-
f
U|«jirl PC, Fe(0) Dwngrd. PG Dwngrd.
DAEtyr
' PG Fs(0) D^vnyr ' PG
Figure 4.12 Average (± 1 s.d) concentrations of Na, K, Ca, Mg, sulfate, bicarbonate, chloride, and silica (mg/L) as a
function of sampling position in gate 1 at the Denver Federal Center.
65
-------
DFC 2
250-
200-
150-
100-
50-
0-
I
1
f
H'H
0 9
^ °-6-
0.3-
00
I
1
—
-
Upgrd. Upgrd. PG Fe{0) Dwngrd. PG Dwngrd.
Upgrd. Upgrd. PG Fe(0) Dwngrd. PG Dwngrd.
Upgrd. Upgrd. PG Fe{0) Dwngrd. PG Dwngrd.
Upgrd. Upgrd. PG Fe(0) Dwngrd. PG Dwngrd.
y/y*
y,y>
/y/y
y,y,
s/s/.
'X'X
////
F
Upgrd. Upgrd. PG
y,y>
PP
m
/ ////
7//A
y,y>
Dwngrd. PG Dwngrd.
Upgrd. Upgrd. PG P®{0) Dwngrd. PG Dwngrd.
18-,
16-
14-
12-
ti>^
Upgrd. Upgrd. PG
Dwngrd. PG Dwngrd.
Upqid Ufqid PG Fe(0) Dwngrd. PG Dwngrd.
Figure 4.13 Average (± 1 s.d.) concentrations of Na, K, Ca, Mg, sulfate, bicarbonate, chloride, and silica (mg/L) as a
function of sampling position in gate 2 at the Denver Federal Center.
66
-------
DFC Gate 3
300-.
250-
150-
100-
50-
0
I
1
r+n
i
i
1
I
I
i
i
3.5-.
3.0-
2.5-
2.0-
=d
if
1.0-
0.5-
0.0
,1
I
1
I
+
T
i
Upgrd. Upgrd. PG Fe{0) Dwngrd. PG Dwngr
Upgrd. Upgrd. PG Fe{0) Dwngrd. PG Dwngrd.
280-,
240-
200-
160-
50-,
45-
40-
35-
30-
25-
20-
15-
10-
5-
0-
Upgrd. Upgrd. PG Fe{0) Dwngrd. PG Dwngrd.
Upgrd. Upgrd. PG Fe(0) Dwngrd. PG Dwngrd.
500 ,
450-
400-
350-
300-
250-
200-
150-
100-
50-
0--
Upgrd. Upgrd. PG Fe{0) Dwngrd. PG Dwngrd.
Upgrd. Upgrd. PG Fe{0) Dwngrd. PG Dwngrd.
Upgrd. Upgrd. PG
Dwngrd. PG Dwngrd.
Upgrd. Upgrd. PG Fe{0) Dwngrd. PG Dwngrd.
Figure 4.14 Average (± 1 s.d.) concentrations of Na, K, Ca; Mg; sulfate, bicarbonate, chloride, and silica (mg/L) as a
function of sampling position in gate 3 at the Denver Federal Center.
67
-------
In downgradient wells (GSA-20, GSA-25, and GSA-30), concentrations of sodium, calcium, magnesium, and potassium
were broadly comparable to those measured in upgradient-monitoring locations (Figures 4.12-4.14). Calcium concentra-
tions are slightly depleted (30% reduction) in downgradient well GSA-25 (gate 2) compared to Ca concentrations in
upgradient well GSA-26, In well GSA-30, located downgradient from gate 3, concentrations of sodium and calcium are
greater than those in upgradient well GSA-31. As noted previously, specific conductance values are also greater in well
GSA-31 compared to well GSA-30 indicating that gate 3 only partially impacts the chemistry of downgradient ground
water. Iron concentrations in downgradient monitoring points have always been below 0.5 mg/L (<0.01 mM).
Interestingly, influent ground water to gate 3 is elevated in ferrous iron (>15 mg/L), perhaps due to an influence of anoxic
bottom waters in Downing Reservoir. Dissolved iron concentrations, however, are expected to decrease within the
reactive media due to the corrosion-induced pH increase, consequent development of oversaturated conditions with
respect to iron hydroxide precipitates, and iron partitioning into the solid phase.
In contrast to the cationic compositions of ground water collected from upgradient wells at the DFC that show only a
moderate amount of variability, anionic compositions are more variable. In gates 1-3, the predominant anion on molar
basis is bicarbonate, which ranges in concentration from about 5.9 mM (gate 2) to 11.9 mM (gate 3). Sulfate
concentrations are fairly consistent in upgradient wells from gate 1 (2.5 mM) and gate 2 (2.9 mM), but sulfate was not
detected in ground water upgradient of gate 3 (<0.01 mM). Chloride concentrations are fairly uniform, ranging from
1.5 mM upgradient of gate 1 to 2.1 mM upgradient of gate 3. Concentrations of nitrate entering the funnel-and-gate
system at the DFC are low to below detection limits (<0.03 mM). Alkalinity concentrations are between 26% and 55%
lower in the reactive cells than in upgradient sampling wells. Sulfate concentrations are often below detection limits
within the reactive cells; however, higher concentrations of sulfate (relative to upgradient points) were observed in wells
located within the reactive media of gate 2 in 1999 and 2000 (up to 154% higher). This trend in sulfate concentrations
in gate 2 is highly unusual compared to other PRB systems and suggests that some re-oxidation of sulfide precipitates
was occurring, perhaps due to infiltration of oxidizing ground water. In 2001, however, the trend in sulfate concentrations
in gate 2 reversed, i.e., lower concentrations were detected within the reactive media, but not to the >95% depletion
levels typical in gate 1 or at the Elizabeth City PRB as previously described. Nitrate was never detected in mid-wall
monitoring wells or in downgradient wells.
Chloride concentrations are significantly higher in downgradient wells compared to upgradient wells. Up to about two
times as much chloride is found in downgradient wells GSA-20, GSA-25, and GSA-30 as compared to upgradient wells
GSA-21, GSA-26, GSA-31 (Figures 4.12-4.14). Similarly, sulfate concentrations are greater in the downgradient regions
of gate 1 and gate 3, whereas concentrations of bicarbonate are lower. In general, the geochemical trends at the DFC
are fairly atypical of other PRB sites investigated in the Tri-Agency initiative. Well transects across the various reactive
cells do not appear to show geochemical connectivity between upgradient and downgradient regions. As was observed
in the contaminant distributions, the geochemistry of downgradient ground water is only partially represented by the
chemistries of ground water emerging from the various treatment cells.
4.2 at the
Core samples were collected at the DFC from gate 1, gate 2, and gate 3 in July 2000 and 2001. Similar core collection
methods to those used at the Elizabeth City site were adopted at the DFC. Core collection methods and analysis
procedures are described in Volume 2 of this EPA Report series. In all cases 5 cm inner diameter cores were collected
using direct-push methods (Geoprobe™). Angled cores (30° relative to vertical) and vertical cores were collected in
order assess the spatial distribution of mineral/biomass buildup in the reactive media. Prior to pushing the core barrel,
an electrical conductivity profile was collected to verify the exact position of the iron/pea gravel interface.
Core materials from the DFC PRB were jet black in color. Cores collected in July 2000 showed no obvious signs of
cementation. Several cores collected in July 2001, however, contained sub-spherical welded nodules of iron grains up
to about 3cm in diameter (Figure 4.15). Iron grains from the upgradient interface of DFC gate 2 were noticeably enriched
in a black-colored, gel-like material (mixture of biomass and fine-grained mineral precipitates). This core consistency
was not observed at other DFC or at the Elizabeth City PRB (Figure 4.16).
Immediately after collection, the cores were frozen and shipped back to the Ground Water and Ecosystems Restoration
Division in Ada, OK for sub-sampling and analysis. The frozen cores were partially thawed and then placed in an
anaerobic chamber maintained with a 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 4 sub-samples: (1) inorganic
carbon analyses, (2) sulfur analyses/X-ray diffraction (XRD), (3) Scanning electron microscopy (SEM)/X-ray photoelec-
tron spectroscopy (XPS) analyses, and, (4) microbial assays (phospholipid fatty acids, PLFA). All sub-samples were
retained in airtight vials to prevent any air oxidation of redox-sensitive constituents. Details of analytical methods used
to characterize the core materials are presented in Volume 2 of this EPA Report series.
Locations of coring events at the DFC are shown in Figures 4.1-4.3 and information about core recovery, core length,
and depth of core penetration is presented in Table 4.1. Cores collected from the upgradient aquifer/iron region generally
penetrated the PRB at depths of 4.5 to 6.5 m below ground surface.
68
-------
Figure 4.15 Picture showing cemented nodules recovered from a gate 1 core collected at the Denver Federal Center
(cored-1-71701).
Figure 4.16 Picture showing the appearance of a core collected at the Denver Federal Center, from gate 2 near the
upgradient pea gravel/iron interface.
69
-------
Table 4.1. Cores collected for analysis at the Denver Federal Center PRB
-vl
o
Sample ID
C1 -1-71 701
C1 -2-71 701
C1 -3-71 701
C1 -4-71 801
C2-1-71801
C2-3-71801
C2-4-71801
C2-1-71901
C2-2-71901
C2-3-71901
C3-2-71801
C1 -2-71 000
C1 -3-71 100
C2-1 7-71 300
C2-1 6-71 300
C2-1 2-71 300
C2-1 3-71 300
C2-1 4-71 300
Location
Upgradient
Upgradient
Upgradient
Downgradient
Upgradient
Upgradient
Upgradient
Vertical
Vertical
Vertical
Upgradient
Upgradient
Downgradient
Upgradient
Downgradient
Vertical
Vertical
Vertical
Date
July-01
July-01
July-01
July-01
July-01
July-01
July-01
July-01
July-01
July-01
July-01
July-00
July-00
July-00
July-00
July-00
July-00
July-00
Angle
30
30
30
30
30
30
30
90
90
90
30
30
30
30
30
90
90
90
Core length
cm
102
79
81
112
117
97
97
114
107
91
81
82
81
43
28
102
91
71
Recovery
%
83
65
67
92
96
79
79
94
88
69
67
73
69
71
46
83
75
62
Depth to Fe
m bgs
5.3
6.3
5.3
5.3
5.8
5.8
6.1
4.2
4.2
4.2
5.5
4.8
5.3
5.8
4.3
4.2
4.2
4.2
Notes: All gate 2 and gate 3 cores (C2, C3) were collected by pushing the core barrel from the downgradient side of the reactive barrier. Depth to iron is the depth
below ground surface in meters at which zero-valent iron was intercepted. The percent recovery was calculated as (length compacted core/push Iength)x100.
-------
4.2.1
Results of all carbon analyses of DFC core materials (n = 251) are presented in Table A3 (Appendix A). Concentrations
of inorganic carbon within >95% iron samples range from about 10 to 11,600 |ig/g. In all cases the highest
concentrations were found in samples collected adjacent to the upgradient or downgradient pea gravel/iron interfaces
and the lowest concentrations were detected near the center regions of the iron gates.
The concentration distribution of inorganic carbon in cores collected from DFC gate 1 is shown in Figure 4.17. The
maximum concentration near the upgradient edge of gate 1 was observed in core C1-1-71701 (8700 p.g/g) in material
that contained a mixture of Fe° and pea gravel (Figure 4.17). A concretion was recovered in this core at the location
corresponding to the maximum carbonate concentration (Figure 4.17). Inorganic carbon concentrations fell below
300 jig/g in gate 1 at horizontal penetration depths >25 cm. Core materials were not collected from the mid-barrier
regions, but the trends shown in Figure 4.17 suggest that carbonate accumulation near the center of the iron wall is
negligible. Surprisingly, elevated inorganic carbon concentrations were observed in two cores collected at the
downgradient pea graveI/Fe° interface region (Figure 4.17). Concentrations in core C1-4-71801 were even higher than
those detected near the upgradient region of the reactive media. Concretions up to 2-cm in diameter were recovered in
samples that contained between about 8,000 and 11,600 ja,g/g inorganic carbon.
Gate 2 shows trends in inorganic carbon concentrations that are similar to those observed in gate 1. Concentrations as
high as 7,700 fig/g were detected incoreC2-3-71801; concretions were also identified in this core (Figure 4.18). Vertical
cores were collected in gate 2 (C2-1 -71901, C2-2-71901, and C2-3-71901) to examine the depth-dependent distribution
of carbon and sulfur accumulation in the reactive media. The vertical sampling position was located approximately 15 cm
downgradient of the upgradient pea gravel-Fe0 interface (Figure 4.2). Results show that there is a depth interval between
about 5.5 m and 8.0 m below ground surface where very little inorganic carbon accumulation is occurring (Figure 4.19);
significant amounts of carbonate precipitation has occurred only near the bottom and the top of the reactive zone. The
implication of this trend, which is also mirrored in the concentration profiles of sulfur and microbial biomass, is that little
ground water is entering gate 2 over the depth interval from about 5.5 m to 8 m below ground surface. Flow would appear
to be occurring near the very top and the bottom of the reactive cell. This result could help explain the anomalous
behavior of this gate with respect to contaminant removal performance.
4.2.2
Results of sulfur analyses of DFC core materials strongly correlate with the inorganic carbon results (Figures 4.20-4.21).
Concentrations of sulfur in core materials from the DFC range from about 100 U-.g/g to 7,500 jig/g (Table A3). Chemical
extractions indicate that over 90% of the total sulfur present in the core materials is as sulfide, in acid-volatile sulfide
materials such as poorly crystalline to crystalline mackinawite (Table A2; Wilkin et al., 2003). The remaining sulfur is
likely present as iron disulfides (pyrite) and perhaps as sulfate associated with iron corrosion products (Furukawa et al.,
2002). Figure 4.22 shows the good correlation between concentrations of carbon and sulfur in the solid phase in cores
collected from the Elizabeth City and Denver Federal Center PRBs. The average S/C ratio in core materials from the
DFC is about 0.42, which is slightly greater than the S/C ratio observed in Elizabeth City core materials (S/C=0.33). The
difference is apparently related to a higher average ground-water S/C ratio at the Denver Federal Center site compared
to the Elizabeth City site.
4.2.3 X-ray
Powder X-ray diffraction scans for samples from DFC core C2-3-71801, collected in 2001, are shown in Figure 4.23.
Materials for analysis were obtained by sonicating iron core samples in acetone for 10 minutes followed by collection of
the released particulates on a 0.2-micron filter paper (polycarbonate). The separated particles were then mounted on a
zero-background quartz plate and scanned with Cu Ka radiation from 3° to 80° 2-theta using a Rigaku Miniflex
Diffractometer.
A summary of the XRD analysis results is reported in Table 4.2. Qualitative abundances of the mineral phases identified
are reported based upon observed peak intensities (Table 4.2). Magnetite (Fe3O4) and iron carbonate hydroxide were
observed in every iron core sample. Graphite and quartz were detected in one of the samples (C2-3-71801-3).
Aragonite (CaCO3) was not detected in the diffraction analysis, although the presence of calcite was confirmed using
XRD, SEM and optical microscopy. Using micro-analytical diffraction methods, Furukawa et al. (2002) also detected
quantities of mackinawite, greigite, ferrihydrite, and goethite in core materials collected from the DFC.
SEM photomicrographs for three samples from the DFC collected in July 2000 are shown in Figure 4.24. These samples
were retrieved from near the upgradient pea gravel/iron interface region from gate 2 (C2-17-71300-2, horizontal
penetration ~4 cm; Figure 4.24a, C2-17-71300-7, horizontal penetration ~16 cm; Figure 4.24b) and gate 1 (C1-2-71000-3,
horizontal penetration ~ 6 cm; Figure 4.24c). The SEM micrographs in Figure 4.24 are representative of particles
contained in each of the samples and capture a range of magnifications from about 50x to 2000x.
71
-------
-5?
O)
0
_Q
S—
CD
O
0
'c
CD
E?
0
c
14000-
-
12000-
10000-
8000-
•
6000-
4000-
2000-
0-
0 -«-C1
2 -*-C1
« -A-C1
a -?-ci
-1-71701
-2-71000
-3-71100
-4-71801
i /
co Fe(0) + pea gravel Concretions observed "
O) •
Q. '
Concretions observed
>>
I
• « > t « <»W'
"c
_0
1 ~O
I CD
ft D)
C
1
IT
r :
: T
\i
0
T3
'
^v
L4A: T
0
CD
D)
0
Q.
0
200
50 100 150
Distance, cm
Figure 4.17 Concentration distribution of solid phase inorganic carbon in angle cores collected from gate 1 at the
Denver Federal Center.
8000-
7000-
6000-
- 5000-
o
I 4000-
O
~ 3000-
CD
O)
fe 2000-
1000-
0-
15 30
Distance, cm
C2-17-71300
C2-3-71801
C2-4-71801
45
60
Figure 4.18 Concentration distribution of solid phase inorganic carbon in angle cores collected from gate 2 at the
Denver Federal Center.
72
-------
5-
6-
E
(/) 7-
O
CO
| 8-
9-
10-
•-^
T
i C=~
- ^^
^^\
•i*^.
— •— C2-[1-3]-71901
•
."-•^.
0 2000 4000 6000
8000
Inorganic Carbon, (ig/g
Figure 4.19 Concentratio
Denver Fede
800(
700(
n distribution of solid phase inorganic carbon in a vertical core collected from gate 2 at the
ral Center.
)-
6000-
.o> 5000-
Ł 4000-
w 3000-
_co
B
!~ 2000-
1000-
0-
; — •— C1-1-71701
; — •— ci -2-71 ooo
i --A-C1 -3-711 00
; — ?— C1-4-71801
^
: \
fifk
I -fc^
J
t
A
AT :
/V :
1 \
'
•A
:/
1
4
-i— «
o
CO ^
L— (^5
C O)
0 CD
T3 Q.
L
r
r
V
0
50 100 150
Distance, cm
200
Figure 4.20 Concentration distribution of solid phase sulfur in angle cores collected from gate 1 at the Denver
Federal Center.
73
-------
s_r
^
4000-
3000-
2000-
1000-
0-
\
\
C2-17-71300
C2-3-71801
C2-4-71801
•
\
W
15 30 45
Distance, cm
60
Figure 4.21 Concentration distribution of solid phase sulfur in angle cores collected from gate 2 at the Denver
Federal Center.
Sample C2-17-71300-2, collected from a region near the upgradient interface in gate 2, shows a fairly regular
accumulation of mineral precipitates on the iron surfaces (Figure 4.24a). The thickness of the surface coating ranges
from about 10 u,m to 100 u,m. This precipitate accumulation took place over the initial four years of operation of the PRB,
or at a rate of about 3 to 25 u,m per year, similar to the average linear rate of precipitate accumulation observed at the
Elizabeth City site. As noted before, ground-water chemistry is different between these two sites as is the total mass of
mineral precipitate accumulation. Rates of inorganic carbon and sulfur accumulation are between 2 and 10 times greater
within the DFC reactive media as compared to the Elizabeth City reactive media (Wilkin et al., 2003). The similar
average thickness of mineral precipitates found on iron grains near the upgradient edge at the two sites suggests that
mineral precipitation at the DFC must occur over a wider range of penetration depths, i.e., with time a mineral precipitate
front moves through the reactive barrier.
The DFC samples contain a significant proportion of free grains, grains not directly bound to the iron surfaces
(Figure 4.24a). In all cases, these large free grains (up to 300 u,m in diameter) are composed of calcium carbonate. At
greater horizontal penetration depths in gate 2, surface coatings were thinner, 10 to 50 u,m, but the abundance of free
calcium carbonate grains persisted (Figure 4.24b). It cannot be determined conclusively whether the free grains actually
grew within pore spaces, or whether they were at one point attached to the iron surfaces, but were subsequently
detached, for example, during vibratory coring or during sample handling. In gate 2, surface coverage and particle
morphologies are similar to those observed in gate 1 (Figure 4.24c). The micrographs clearly reveal an abundance of
calcium carbonate grains associated with the iron surfaces.
Energy dispersive X-ray analyses of iron grains from the DFC (Table 4.3) indicate the presence of iron (~97 wt%), silicon
(~2.3%), and to some extent Mn (<1.1 wt%) and Cr (<1.2 wt%). Oxygen was not detected on freshly polished surfaces
of zero-valent iron. The compositions of Ca- and O-rich grains (presumably calcium carbonate) associated with iron
surfaces are identical to those found as free grains (Table 4.3). These particles appear to be enriched in calcium
compared to ideal CaCO3 (40.0 wt% Ca), and this is probably related to quantitative inaccuracies of the EDX method.
Surface precipitates, i.e., materials that are generally fine-grained and coat the iron surfaces, are enriched in oxygen and
depleted in iron, respectively, compared to the composition of fresh zero-valent iron (Table 4.3). In addition to iron and
74
-------
4000-
3500-
3000-
~§> 250°-
13
^- 2000-
^ 1500-
-§ 1000-
a) Elizabeth City cores
3 = 128.6 + 0.33*10
R2 = 0.48, n=170
1000
2000 3000 4000 5000
Inorganic carbon, ug/g
6000
I
en
1
10000-,
9000-
8000-
7000-
6000-
5000-
4000-
3000-
2000-
1000-
0-
-1000-
b) Denver Federal Center cores
3 = 712.9 + 0.42*10
R = 0.62, n=251
-2000 0
4000 8000 12000
Inorganic Carbon, ug/g
16000 20000
Figure 4.22 Inorganic carbon concentrations versus total sulfur: a) Elizabeth City core materials; b) Denver Federal
Center core materials.
75
-------
3500-
3000-
2500-
2000-\
CO
Ł 1500-
1000-
500-
0
DFC core C2-3-71801
mid-barrier
-7
upgradient edge
0 10 20 30 40 50 60 70 80 90
degrees 2-theta
Figure 4.23 Powder X-ray diffraction data from fine-grained materials removed via sonication from cores collected at
the Denver Federal Center PRB (core C2-3-71801).
Table 4.2. Results of Powder X-ray Diffraction Analysis of Core Materials from the Denver Federal Center PRB
Sample
C2-3-71 801-1
C2-3-71801-2
C2-3-71801-3
C2-3-71801-5
C2-3-71801-7
C2-3-71801-15
Major Component
Magnetite, Iron
carbonate hydroxide
Magnetite, Iron
carbonate hydroxide
Magnetite, Iron
carbonate hydroxide
Magnetite, Iron
carbonate hydroxide
Magnetite, Iron
carbonate hydroxide
Magnetite, Iron
carbonate hydroxide
Minor Component
Green rust 1 , Calcite
Mackinawite, Calcite
Mackinawite, Calcite
Mackinawite, Calcite
Mackinawite, Calcite
Calcite
Trace Component
Goethite
Quartz, Graphite
76
-------
15kV 100x 100iim
15kV 500x SOiim
15 kV 350x SOiim
Figure 4.24 Scanning electron micrographs of samples from the Denver Federal Center PRB:a) sample C2-17-
71300-2 located near the upgradient edge of gate 2; b) sample C2-17-71300-7 located in a midbarrier
region of gate 2; and, c) sample C1 -2-71000-3 located near the upgradient edge of gate 1.
77
-------
Figure 4.24 (continued) Scanning electron micrographs of samples from the Denver Federal Center PRB:a) sample
C2-17-71300-2 located near the upgradient edge of gate 2; b) sample C2-17-71300-7 located in a
midbarrier region of gate 2; and, c) sample C1-2-71000-3 located near the upgradient edge of gate 1.
78
-------
15kV 350x
'»
15kV 1500x 10iim
Figure 4.24 (continued) Scanning electron micrographs of samples from the Denver Federal Center PRB:a) sample
C2-17-71300-2 located near the upgradient edge of gate 2; b) sample C2-17-71300-7 located in a
midbarrier region of gate 2; and, c) sample C1-2-71000-3 located near the upgradient edge of gate 1.
79
-------
Table 4.3. Results of SEM-EDX Analysis of Core Materials from the Denver Federal Center PRB
Iron
(n=25)
Ca-rich
precipitates,
on iron
surfaces
(n=23)
Ca-rich
precipitates,
free
Surface
precipitates
(n=57)
Fe
(wt%)
97.4
(±0.7)
0.2
not
detected
58.3
(±10.5)
O
(wt%)
not
detected
54.7
(±3.5)
54.8
(±1.5)
31.6
(±5.5)
S
(wt%)
not
detected
not
detected
not
detected
3.9
(±3.0)
Mn
(wt%)
0.15
(±0.14)
not
detected
not
detected
0.34
(±0.7)
Cr
(wt%)
0.3
not
detected
not
detected
0 04
Ca
(wt%)
not
detected
44.5
(±3.4)
44.6
(±1.9)
1.2
(±5.2)
Mg
(wt%)
not
detected
not
detected
not
detected
0.11
(±0.3)
Si
(wt%)
2.30
(±0.4)
not
detected
not
detected
3.89
(±3.0)
oxygen, the surface precipitates also contain silicon (3.4 ± 1.9 wt%, n=58), sulfur (3.89 ± 3.0 wt%, n=58), manganese
(0.34 ± 0.70 wt%, n=58), chromium (0.04 ± 0.2 wt%, n=58), calcium (1.2 ± 5.2 wt%, n=58), and magnesium
(0.11 ± 0.30 wt%, n=25). The presence of Ca, Mg, and Si in the surface coatings is consistent with their removal from
ground water. Other than CaCO3, the identity of mineral species could not be confirmed by using EDX measurements.
42.5
Eighty-one samples from 2000 to 2001 were collected at the DFC for phospholipid fatty acid (PLFA) extract
characterization. Samples were collected mainly from within the reactive iron media near the upgradient iron/pea gravel
interface and near the downgradient iron/pea gravel interface. Samples for PLFA analysis were shipped frozen on dry
ice to Microbial Insights (Rockford, Tennessee). The complete PLFA data set for samples from the DFC is shown in
Table B2 (Appendix B) and summarized in Table 4.4.
In samples collected at the DFC, biomass contents spanned several orders of magnitude from 15 to 4924 picomoles per
gram (dry weight basis), or from about 3.85x105 to 9.85x107 cells per gram. The highest biomass contents observed in
this study were in gate 2 of the DFC, where maximum PLFA concentrations were about 1.5 to 3 times the maximum
concentrations observed in DFC gate 1 or in the Elizabeth City PRB.
In gate 1 and gate 2, the highest biomass concentrations were found near the upgradient gravel/iron interface region,
in the same portion of the reactive media where enrichments in inorganic precipitates are observed (Figure 4.25). This
same trend was observed at the Elizabeth City PRB and appears to be repeated at other PRB sites (Gu et al., 2002).
Downgradient regions are comparatively depleted in microbial biomass. The lower counts associated with the mid-
barrier and downgradient samples suggest that the environment at these locations is more challenging to bacterial
growth and survival. Vertically resolved biomass concentrations and total sulfur concentrations from gate 2 are plotted
in Figure 4.26. The results parallel those previously shown for inorganic carbon (Figure 4.19). The depth interval
between about 5.5 and 8.0 m below ground surface is characterized by low accumulation of sulfur and microbial biomass
suggesting stagnant conditions in this portion of gate 2.
PLFA profiles from DFC are typically dominated by fatty acid biomarkers indicative of anaerobic sulfate- or iron-reducing
bacteria (Figure 4.27). High proportions of terminally branched and branched monoenoic PLFA specifically indicate
anaerobic metabolism. Terminally branched PLFA are typical of Gram-positive bacteria, but can also be present in the
cell membranes of some anaerobic Gram-negative bacteria. Because high proportions are present of branched
80
-------
Table 4.4. Summary of PLFA Data from the Denver Federal Center PRB
Number of samples
Average PLFA
concentration
(pmoles/g)
PLFA range
PLFA Structural
Groups
average % (range %)
Monoenoic
Found in Gram-negative
bacteria
Terminally Branched
Saturated
Found in many Gram-
positive bacteria, and in
some Gram-negative
bacteria
Branched Monoenoic
Common in obligate
anaerobes, such as
sulfate-reducing and
iron-reducing bacteria
Mid-Chain Branched
Saturated
Common in
actinomycetes and
sulfate-reducing bacteria
Normal Saturated
Ubiquitous in
prokaryotes and
eukaryotes
Polyenoic
Found in fungi,
protozoa, algae, higher
plants, and animals
Fe(0)
Upgradient
DFC - Gate 1
9
841
95-1904
59.6
(10.1-71.2)
7.0
(2.7-11.4)
8.3
(2.0-11.1)
4.2
(1.7-6.2)
18.5
(4.6-23.2)
2.4
(0.2-6.7)
Fe(0)
Downgradient
DFC - Gate 1
9
49
15-127
77.4
(9.2-87.8)
2.1
(<1-6.9)
3.7
(<1-8.9)
1.3
(<1-4.7)
12.2
(3.8-17.3)
3.3
(<1-9.4)
Fe(0)
Upgradient
DFC - Gate 2
38
1440
50-4924
52.2
(31.6-85.3)
14.2
(3.4-26.4)
10.1
(<1-16.5)
6.1
(<1-22.2)
15.4
(8.9-30.6)
2.0
(<1-8.2)
Fe(0)
Downgradient
DFC - Gate 2
6
98
24-332
69.6
(54.9-77.7)
6.8
(3.1-26.4)
2.6
(<1-8.4)
2.4
(<1-10.6)
11.2
(8.6-13.2)
7.4
(1.7-18.4)
81
-------
2500-
0) 2000-
o5
Q
§ 1500-
Lij 1000-
Q.
500-
0-
a;
n
— i
"
i — i
^^B Gate 1, Fe(0) upgradient
i i Gate 1, Fe(0) downgradient
n™ ,
1
ru-
34567
Ranked Sample Distribution
5000-
4000-
3000-
2000-
1000-
b)
Gate 2, Fe(0) upgradient
Gate 2, Fe(0) downgradient
10 15 20 25 30
Ranked Sample Distribution
35
Figure 4.25 Ranked concentration distribution of microbial biomass (from Denver Federal Center PLFA data) in
picomoles per gram in iron from near the upgradient pea gravel/iron interface and iron from near the
downgradient pea gravel/iron interface: a) gate 1; b) gate 2.
monoenoic PLFA indicative of anaerobic metal reducing bacteria, the terminally branched PLFA are likely to be mainly
from sulfate- or iron-reducing bacteria. Where biomass is most concentrated (i.e., near the upgradient pea gravel/iron
interface), the distribution of PLFA overall appears to be distinct from the PLFA distribution observed in regions further
downgradient (Figure 4.27). Near the upgradient pea gravel/iron interface, the proportion is greater of branched
monoenoic PLFA and PLFA indicative of sulfate-reducing bacteria compared to the PLFA signature of materials
collected at the downgradient pea gravel/iron interface (Figure 4.27).
4.3 Summary of Results from the Denver Federal Center Site
The Denver Federal Center permeable reactive barrier is a funnel-and-gate system with four reactive gates, each
separated by up to about 120 m of metal sheet pile. In this study, ground-water sampling, core collection, and solid
phase characterization studies were carried out in gates 1, 2, and 3. After five years of operation, gate 1 and gate 3 have
been effective in significantly decreasing concentrations of TCE, 1,1,1-TCA, 1,1-DCE, and cis-DCEfrom influent ground
water. A noticeable clean ground-waterfront is identifiable in the downgradient aquifer near gate 1. The performance of
gate 2 has been more difficult to assess, in part, because of an apparent source of residual bedrock TCE contamination
82
-------
5-
6-
CO"
o
00
±j-
&
8-
9-
10-
DFC Gate 2
Vertical Core
Total Sulfur,|ig/g
Biomass (PLFA), picomoles/g
0 500 1000 1500 2000 2500 3000 3500 4000
Total Sulfur, Biomass
Figure 4.26 Concentration distribution of solid phase sulfur and microbial biomass (from PLFA data) in a vertical core
collected from gate 2 at the Denver Federal Center (vertical cores C2-1 -71901, C2-2-71901, and C2-3-
71901).
located just downgradient of gate 2 and because of the large hydraulic head differential across the gate. Concentrations
of volatile organic compounds in gate 2, especially 1,1-DCE, have shown increases with time. Results of the Denver
Federal Center site study are summarized below:
• Trends in pH in gates 1, 2, and 3 were fairly consistent from May 1999 to July 2001. There is some
indication that pH in the iron zone is slowly increasing with time, perhaps due to increasing residence time.
Measurements of Eh and hydrogen gas concentrations indicate consistent reducing conditions in gates 1
and 3. Gate 2 shows evidence of having decreased reducing potential, based on increased Eh and
decreasing hydrogen gas concentrations.
• The reactive gates at the Denver Federal Center have removed most of the dissolved calcium, magnesium,
sulfate, and silicon in the water flowing through the PRB. Levels of alkalinity and total dissolved solids were
reduced. Ground water from the DFC is comparatively (~3x) more enriched in concentrations of total
dissolved solids compared to ground water from the Elizabeth City site. Consequently, more mineral
precipitate mass has accumulated over five years in the DFC gates compared to the Elizabeth City PRB.
After five years, core sampling revealed the presence of cemented iron nodules in some of the cores
collected at the DFC.
• Mineralogical characterization of soil cores indicates the formation in the iron gates of calcite, iron carbonate
hydroxide, magnetite, mackinawite, carbonate green rust, and goethite. Overall, this assemblage of mineral
precipitates is very similar to that observed at the Elizabeth City site and at other PRB installations. Mineral
83
-------
8.3%
4.2%
19%
60%
77%
12%
3.3%
2.1%
Gate 1 upgradient
Gate 1 downgradient
52%
10%
6.1%
15%
70%
2.6%
2.4%
11%
Gate 2 upgradient
Gate 2 downgradient
I Gram+/Gram- (Ter. Br. Sat.)
] Gram- (Monoenoic)
I Anaerobic (Br. Mono.)
] SRB, Actinomycetes
I Genera Nsats
I Polyenoic
Figure 4.27 Pie graphs showing average structural distribution of PLFA compounds in Iron core materials from the
Denver Federal Center.
84
-------
precipitates are enriched in gate 1 and gate 2 at both the upgradient and downgradient edges of the reactive
medium. After five years of operation, up to 50% of the available pore space has been lost due to infilling of
mineral precipitates near the leading edges of the iron treatment zones. The high infilling rate has led to
partial cementation of the iron zone. Because of the high influent TDS concentrations and rates of
mineralization, hydraulic performance of the system is expected to degrade over the next five years.
Vertical coring in gate 2 reveals a zone from 5.5 to 8 m below ground surface with little mineral or microbial
biomass accumulation, suggesting a substantial zone of reduced flow through the PRB. This reduced flow
zone may be related to the low permeability zone or smear zone that developed during construction of the
gate.
Microbiological results, based on PLFA analyses, indicate enriched biomass in gate 2 of the Denver Federal
Center (up to 9.85x107 cell/g). Biomass concentrations in gate 2 were a factor of about 2x greater compared
to 1 and 3, and the Elizabeth City PRB. The higher biomass concentrations may be linked to low-flow
conditions in the gate and high sulfate concentrations. Concentrations of iron sulfide precipitates are also
comparatively high in gate 2 of the DFC (up to 5650 pg/g). Gram-negative bacteria dominate the microbial
community. Present in the iron samples are elevated proportions of biomarkers indicative of metal-reducing
and sulfate-reducing bacteria.
85
-------
86
-------
5.0
Permeable Reactive Barriers are principally designed to remove contaminants from ground water, yet it has been shown
that zero-valent iron reactive barriers are long-term sinks for other dissolved components in ground water. In particular,
inorganic carbon, sulfate, calcium, magnesium, nitrate, and silica are either entirely or partially removed from influent
waters to PRBs. Removal processes for inorganic constituents include mineral precipitation, adsorption, and biochemi-
cal transformation. Mineral precipitation, in particular, can lead to changes in porosity and permeability, or both, in
addition to affecting the reactivity of zero-valent iron media by forming coatings on the reactive grains. In this way,
geochemical reaction processes can have an effect on the hydraulic and reactive performance of PRB systems. A major
research effort, therefore, has gone into 1) identifying the nature of element removal processes in PRBs, 2) linking these
processes to site geochemistry and hydrogeology, and 3) linking these processes to declining remedial performance.
Considerable progress has been made on the first question relating to identifying element removal process. Observa-
tions from numerous PRB installations and laboratory studies are used below to develop a framework for understanding
how ground-water chemistry may be evaluated for the purpose of predicting longevity and barrier performance through
time. This framework will be useful in constructing remedial designs and in developing performance-monitoring
programs.
Iron metal dissolution, microbial sulfate reduction, microbial nitrate reduction, adsorption, gas production, and mineral
precipitation (oxides, hydroxides, carbonates, and sulfides) are processes that have been recognized as important
geochemical and biogeochemical processes in zero-valent iron walls. The lifetime of an iron wall will essentially be
determined by the extent that these processes impact the reactive barriers' ability to remove contaminants from ground
water. These processes are discussed in the following sections.
5.1 Fe°
The corrosion of zero-valent iron in aqueous environments has been widely studied (e.g., Davies and Burstein, 1980;
Reardon, 1995; Blengino et al., 1995; Gui and Devine, 1994; Odziemkowski et al., 1998). In water, zero-valent iron may
be oxidized to ferrous- or ferric-iron, leading to dissolution and volume loss of the metal. Under oxic conditions, dissolved
oxygen acts as the electron acceptor and can lead to an increase in pH and the production of ferrous-iron and/or ferric-
iron:
O2 + 2H2O = 2Fe2+ + 4OH~ (5.1)
4Fe2+ + O2 + 2H- = 4Fe3+ + 2OH" (5.2)
Ferric-iron is not expected to remain soluble and will precipitate, for example, as ferric hydroxide:
Fe3+ + 3OH- = Fe{OH)3{s) (5.3)
As oxygen is consumed, iron corrosion reactions result in the production of hydrogen.
Fe° + 2H2O = Fe2+ + H2 + 2OH- (5.4)
Both aerobic and anaerobic iron corrosion reactions lead to an increase in pH. Aerobic corrosion is a more rapid
process, as evidenced by the rapid loss of dissolved oxygen in iron-water systems. As iron corrosion proceeds, iron
mineral precipitates form on or near the surface of corroding iron grains, which increases the thickness of an iron oxide
passivation layer already present at the metal surface (e.g., Ritter et al., 2002). Under anaerobic conditions, hydrogen
gas produced through reaction (5.4) may also temporarily passivate the iron surface. In general, low pH and the
presence of oxidants result in more rapid iron corrosion rates. However, many species abundant in ground water can
affect the rate and pathway of iron corrosion. For example, chloride, carbonate, and sulfate can all affect the corrosion
rate of iron metal.
The major consequences of iron dissolution that follow from reactions (5.1)-(5.4) are the production of OH" (pH increase),
decrease in oxidation-reduction potential (Eh decrease), increase in hydrogen concentration, of ferrous iron, and
possible precipitation of sparingly soluble iron precipitates. Laboratory and field studies have shown that pH values in
iron walls are typically >9 and <1 1 . Eh values measured in zero-valent iron systems are oftentimes below -0.50 V.
87
-------
Measured Eh values in PRBs are likely governed by redox reactions involving Fe(0), Fe(ll), and Fe(lll). For example,
reactions that might bracket measured Eh values in iron walls are:
Fe(OH), + 3H+
~ = Fe° + 3HO
and,
3H2O
(5.5)
(5.6)
Fe(OH)3 + 3H+ + e~ = Fe2
The Eh-pH equation for the ferrihydrite-zero-valent iron couple (P=0.053 V) is:
Eh = P + aRT/ nF log[H+]3 (5.7)
where R is the gas constant, Tis temperature, Fis Faraday's constant, and n is the number of electrons in the balanced
half-cell reaction. Using a AGf° for Fe(OH)3 of -696.3 kJ/mol (Langmuir, 1 997) and AGf° for H2O(I) of -237.18 kJ/mol, we
calculate P = -AG°/nF= 0.053V, so that equation (5.7) reduces to:
Eh = 0.053 - 0.0592 pH
Similarly, the Eh-pH equation for the ferrihydrite-ferrous iron couple is
Eh = 0.975 + 0.0592 log [H+]3 [Fe2+]'1
Assuming ideal behavior and Fe2+ = 0.001 mg/L, equation (5.9) reduces to:
Eh = 1.43-0.178pH
(5.8)
(5.9)
(5.10)
Figure 5.1 shows equilibrium trends in Eh and pH for the Fe°-Fe(OH)3 couple (eqn. 5.8), the Fe°-Fe3O4 couple, and the
Fe2+-Fe(OH)3 couple (eqn. 5.10) compared to field measurements of ground-water pH and Eh from Elizabeth City
monitoring wells located within the zero-valent iron media. The scattering of measured data points generally falls in
between the two equilibrium trends plotted on Figure 5.1. This observation suggests that the measured Eh values in Fe°
PRBs are not equilibrium potentials resulting from one redox pair, but are likely the result of mixed potentials from
multiple redox reactions. These redox reactions are generally bracketed by the Fe°-Fe(OH)3 couple and the
Fe2+-Fe(OH) couple.
100
0-
-100-
-200-
-300-
-400-
-500-
-600-
-700
• ML14, Transect 1 Fe°
• ML24, Transect 2 Fe°
ML34, Transect 3 Fe°
Fe'-Fe(OH)3
Fe°-Fe304
Figure 5.1
8.0 8.5 9.0 9.5 10.0 10.5 11.0 11.5 12.0
pH
Redox -pH diagram showing composition of ground water from the Elizabeth City iron wall compared to
equilibrium trends for the Fe°-Fe(OH)3, Fe°-Fe3O4, and Fe2+-Fe(OH)3couples (FeT=0.001 mg/L).
-------
Eh-pH diagrams showing equilibrium relations in the system Fe-H,O-C-S are shown in Figures 5.2-5.7. These Eh-pH
diagrams were constructed using the EQ3/6 thermodynamic database (Wolery, 1979), modified to include data for green
rust (Drissi et al., 1995; Bourrie et a!., 1999; Retail et a!., 2003), iron sulfides (Benning et a!., 2000), and iron metal. The
diagrams show both predominance areas for aqueous species {dashed lines) and solid phases (bold lines). AH diagrams
were drawn with the activity of iron equal to 10~5. Figures 5.2-5.4 were constructed with no suppressed minerals, i.e.,
they represent the equilibrium description of the systems based on the inclusion of all possible iron phases and aqueous
species in the EQ3/6 database. In some cases, suppressing phases that are not expected to form based upon kinetic
reasoning is useful in constructing and interpreting Eh-pH diagrams. Figures 5.5-5.7 were constructed with selected
phases suppressed as described below.
Equilibrium relations in the system Fe-H9O at 25 °C are shown in Figure 5.2. Within the stability field of liquid water, iron
metal dissolution (corrosion) is expected below pH 8. Note this pH boundary depends on the specified iron activity. At
lower iron activities, the solution-solid boundary will shift to the left so that the Fe2+ field will shrink in size. Above pH 8,
iron metal is expected to corrode and be replaced by FeO and magnetite. At higher redox potentials and over a wide pH
range, hematite (Fe2O3) is a stable mineral. Figure 5.3 shows equilibrium relations in the system Fe-H9O-C (EC = 10~2)
at 25 °C. With carbon in the system, stability fields appear for siderite and carbonate green rust. At the specified
conditions for Figure 5.2, green rust is a stable phase and not metastable as is sometimes assumed, i.e., no mineral
phases were suppressed in order for green rust to appear on the equilibrium diagram. When sulfur is exchanged for
carbon, a broad stability field for pyrite (FeS2) appears at low redox potentials (Figure 5.4).
Note that although there is a strong thermodynamic driving force for forming pyrite, kinetic factors may limit the formation
of pyrite in PRBs. Solid-phase characterization studies indicate that mackinawite (Fe1+ S) is the dominant iron sulfide in
Fe° PRBs with only trace quantities of greigite and pyrite detected (Wilkin et al., 2003;+Furukawa et al., 2002). Because
pyrite is not a major corrosion product in iron walls, suppression of pyrite from the Eh-pH diagram is reasonable (Figure
5.5). As pyrite is removed from the thermodynamic database, other iron sulfides: pyrrhotite, greigite, troilite, and
mackinawite, appear as the stable solid phase at low redox potentials. Figure 5.6 was constructed with all sulfides
suppressed. In this case, magnetite and FeO reappear at pH>8 and low redox potentials, as does a narrow window of
sulfate green rust stability.
Hematite is typically not identified as a corrosion product in iron walls, although Fe(lll)-bearing corrosion products are
common. Suppression of hematite from the thermodynamic database leads to the appearance of goethite (Figure 5.6)
in the Fe-H2O-C system. Further suppression of goethite, magnetite, FeO, and siderite leads to broad fields of carbonate
green rust and ferrihydrite. There is in fact very little kinetic hindrance for the precipitation of these phases (Figure 5.7).
Note that the carbonate green rust field expands as the more insoluble phases hematite and goethite are replaced by
ferrihydrite.
5.2
The anionic composition of ground water is a critical factor in governing the rate of Fe°corrosion and in directing the types
of mineral precipitates that form within Fe° reactive walls. Table 5.1 lists the authigenic minerals that have been identified
in Fe° barriers; these include an assortment of oxides, sulfides, and carbonates. The tendency of ferrous iron to form
complexes with the common anions present in ground water increases in the order CI>HCO3>SO42»OH". Therefore,
it might be reasonably expected that iron metal corrosion rates will be the fastest in chloride-rich ground water and
slowest in bicarbonate- or sulfate-rich ground water. Ground-water chemistry at the Elizabeth City and Denver Federal
Center PRB sites is perhaps typical of many contaminated sites where PRBs might be used in that they contain a mixture
of chloride, bicarbonate, and sulfate (Figure 5.8). If mineral precipitates form as surface coatings on the reactive iron
grains (see Figures 3.24 and 4.24), the effectiveness of iron to degrade halogenated hydrocarbons may be reduced. In
addition, mineral precipitation may result in porosity and permeability reductions (Reardon, 1995). On the other hand,
some mineral coatings may be advantageous for the removal of both organic (e.g., Butler and Hayes, 2000; Lee and
Batchelor, 2002a,b) and inorganic contaminants (e.g., Furukawa et al., 2002), as long as the hydraulic integrity of the
system is retained to prevent ground water bypass of the reactive media.
The amount of mineral precipitation expected in Fe° barriers is linked to the site-specific distribution of anionic species
and to the concentration of total dissolved solids (TDS) in influent waters. High TDS concentrations in influent waters to
Fe° barriers will in most cases result in high mineral accumulation rates. Figure 5.9 shows the range in TDS values
reported from several PRB systems examined in the Tri-Agency research initiative. Note that TDS values are
comparatively low at the Elizabeth City PRB, and this correlates with comparatively low rates of mineral accumulation at
this site.
5.2.1
Inorganic carbon can have an effect on the longevity of zero-valent iron PRBs in two different ways: 1) by accelerating
the corrosion rate of Fe° and the production of dissolved iron and H2, and 2) by forming metal carbonate precipitates
89
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I S I I I I I S I
0
6 8
pH
10
12\ 14
Fe(OH)2'
Figure 5.2 Redox-pH diagram for the Fe-H2O system at 25 aC, showing speciation of soluble iron (dashed lines)
and stability fields of iron-bearing minerals (solid lines). Diagram is drawn assuming iron species
activities of 10~5 (blue shaded area shows field for iron in solution).
02468
Figure 5.3 Redox-pH diagram for the Fe-CO2-H2O system at 25 aC, showing speciation of soluble iron (dashed
lines) and stability fields of iron-bearing minerals (solid lines). Diagram is drawn assuming iron species
activities of 10'5 and carbon species activities of 10'2 (blue shaded area shows field for iron in solution).
90
-------
1.0
0.5
LLJ
-0.5
0
6 8
PH
10
12
14
Figure 5.4 Redox-pH diagram for the Fe-S-CO2-H2O system at 25 °C, showing speciation of soluble iron (dashed
lines) and stability fields of iron-bearing minerals (solid lines). Diagram is drawn assuming iron species
activities of 10'5, carbon species activities of 10'2, and sulfur species activities of 10'3(blue shaded area
shows field for iron in solution).
0
Figure 5.5 Redox-pH diagram for the Fe-S-CO2-H2O system at 25 °C, showing speciation of soluble iron (dashed
lines) and stability fields of iron-bearing minerals (solid lines). Diagram is drawn assuming iron species
activities of 10"5, carbon species activities of 10"2, and sulfur species activities of 10"3(blue shaded area
shows field for iron in solution). Diagram drawn by suppressing all sulfide minerals.
91
-------
Figure 5.6 Redox-pH diagram for the Fe-CO2-H2O system at 25 aC, showing speciation of soluble iron (dashed
lines) and stability fields of iron-bearing minerals (solid lines). Diagram is drawn assuming iron species
activities of 10'5 and carbon species activities of 10~2. Diagram drawn by suppressing hematite and
FeO(s).
V)
"o
LU
i i i i
-0.5 -
10
12
14
2468
pH
Figure 5.7 Redox-pH diagram for the Fe-CO2-H2O system at 25 aC, showing speciation of soluble iron (dashed
lines) and stability fields of iron-bearing minerals (solid lines). Diagram is drawn assuming iron species
activities of 10'5 and carbon species activities of 10'2. Diagram drawn by suppressing hematite, goethite,
magnetite, FeO(s), and siderite.
92
-------
HCO,
Elizabeth City ML10
Denver Federal Center
IDS, mg/L
O 900-1200
o 600-900
0 300-600
° 0-300
Mg2
Figure 5.8 Upgradient ground-water compositions (molar ratios) and IDS values for the Elizabeth City and Denver
Federal Center PRB sites.
Total Dissolved Solids
IDS, mg/L
100
I
Moffett Field (820)
Lowry AFB (2900)
Elizabeth City (250-350)
Denver Fed Ctr (900-1200)
Y-12 (470-3225)
Monticello(1300)
1,000
I
10,000
I
Data source: Tri-Agency PRB Initiative, Combined Report
Figure 5.9 Comparison of total dissolved solids concentrations at PRB sites studied in the Tri-Agency initiative.
93
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which may coat iron metal surfaces, block reactive sites, fill pore space, and through time, impact hydraulic performance.
Abundant field and laboratory evidence indicates that inorganic carbon is removed from ground water during transit
through zero-valent iron PRBs (Vogan et a!., 1999; Phillips et al., 2000; Roh et al., 2000), The decrease in dissolved
inorganic carbon is also accompanied by decreases in the concentrations of calcium and magnesium, Aragonite, calcite,
siderite, ferrous carbonate hydroxide, and the carbonate form of green rust have all been identified in Fe° reactive
barriers (Table 5.1),
Iron corrosion generally leads to a pH increase and a consequent increase in the [CO327HCO3~] ratio in solution.
Generally, influent ground water to iron barriers is saturated to undersaturated with respect to various carbonate
minerals. Increases in pH and the [CO32YHCO3"] ratio will impact reaction affinity and favor precipitation of carbonates
such as aragonite and siderite. At the Elizabeth City PRB, ground water upgradient of multilevel well transect 2 is
undersaturated with respect to both aragonite and calcite (Figure 5.10), based upon the reaction:
Ca2- + HCO - = CaCO3 (ar, cc) + H+
(5.11)
Ground-water compositions from within the iron wall clusters near the saturation point of both aragonite and/or calcite,
and this observation is in agreement with: 1) the measurement of inorganic carbonate in the solid phase, and 2) the
identification of aragonite by powder X-ray diffraction in core materials. Ground water downgradient of the iron wall is
saturated to undersaturated with respect to CaCO3 (Figure 5.10). A similar analysis shows that upgradient ground water
is undersaturated with respect to magnesite and siderite, and that ground water within the Fe° media is near-saturated
to undersaturated with respect to these minerals. Even though ground-water compositions cluster near the saturation
point for Mg and Fe carbonates, we were unable to detect magnesite or siderite in core materials collected from Elizabeth
City. Conclusions about the identity of phases accumulating in reactive barriers that are based solely on the analysis of
mineral saturation states can be misleading. Both magnesite and siderite are known to exhibit very slow
dissolution/precipitation kinetics (e.g., Langmuir, 1997). Solid-phase characterization studies indicate that iron hydroxy
carbonate is, in fact, the dominant mineral carbonate at Elizabeth City, and calcium carbonate and iron hydroxy
carbonate dominate mineral carbonate forms at the Denver Federal Center. There are presently no thermodynamic data
for iron hydroxy carbonate that can be used, for example, to estimate solution saturation indices.
The factors that govern the distribution and form of carbonate precipitates are still unclear. Iron hydroxy carbonate
appears to be a common precipitate documented in both lab and field studies (e.g., this study, Gavaskar et al., 2002;
Kamolpornwijit et al., 2002), The formation of Ca vs. Fe carbonates could in part be controlled by the Ca concentration
of influent water to PRBs. For example, following the exchange reaction:
CaCO. (cc) + Fe2+ = FeCO, + Ca2'
(5.12)
Table 5.1. Mineral Precipitates Identified in Iron Walls
Mineral precipitate type
Oxides and Hydroxides
Carbonates
Sulfides
Minerals identified in
zero-valent iron PRBs
Ferrihydrite, Fe(OH)3
Lepidocrocite, y-FeOOH
Goethite, a-FeOOH
Hematite, Fe2O3
Maghemite, Fe2O3
Green rust 1, Fe6(OH)12CO3-xH2O
Magnetite, Fe3O4
Calcite, CaCO3
Aragonite, CaCO3
Iron carbonate hydroxide, Fe2(OH)2CO3
Siderite, FeCO3
Mackinawite, Fe1+xS
Greigite, Fe3S4
Pyrite, FeS2
94
-------
D)
+
o
D)
O
CD
Li_
+
O
ML24, Transect 2, iron wall
ML21, Transect 2, upgradient
ML25, Transect 2, downgradieni
c)
PH
ML24, Transect 2 - iron wall
ML21, Transect 2 - upgradient
ML25, Transect 2 - downgradient
PH
D
o
+
o
D)
O
• ML24, Transect 2 - iron wall
• ML21, Transect 2 - upgradient
A ML25, Transect 2 - downgradien
a)
o
PH
Figure 5.10 Solubility diagram showing the stability field of a) aragonite, b) siderite, and c) magnesite as a function of
pH and log activities of dissolved inorganic carbon, Ca2+, Fe2+, and Mg2+. Also plotted are ground-water
compositions from upgradient, iron wall, and downgradient sampling locations (Elizabeth City PRB).
95
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with log K „ = 2,2 and ideal behavior in the solid phase, solution concentration ratios of [Ca2+]/[Fe2+]>158 should favor
the formation of calcite over siderite. In batch laboratory tests devoid of calcium, siderite is often found to be the primary
carbonate phase as might be expected (e.g., Agrawal et al., 2002). As the concentration of Ca increases, calcium
carbonates are likely to become more important than siderite. Both aragonite and calcite are observed in iron walls. The
formation of aragonite over calcite may be linked to the [Mg2+/Ca2+] ratio in solution in addition to the absolute Mg2+
concentration. In experiments designed to explore the growth of calcite and aragonite in seawater, Berner (1975)
showed that calcite growth was sensitive to the Mg2+ in solution. In a seawater matrix, the growth rate of calcite was
depressed at Mg2+ concentrations >7Q ppm and aragonite growth was favored from supersaturated solutions. It is also
possible that other ground-water solutes have an effect on directing the nucleation and growth of aragonite over calcite
and subsequent transformation.
Agrawal et al. (2002) recently examined the effect of carbonate precipitation on the reaction between Fe° and
1,1,1-trichloroethane (TCA) in model experiments. Time-dependent trends in degradation half-lives suggest an initial
increase followed by a decrease. The exposure of iron surfaces to bicarbonate solution apparently results in an initial
period of inhibited corrosion due to the presence of a film of iron oxide, followed by bicarbonate-enhanced iron corrosion,
and eventually passivation occurs as a result of carbonate precipitation (siderite and carbonate green rust). These three
regimes yield different contaminant degradation kinetics and perhaps different degradation mechanisms.
5.2.2
Decreases in sulfate concentrations in Fe° barriers are attributed to the microbial activity of sulfate-reducing bacteria. At
both the Elizabeth City and DFC PRBs, sulfate concentrations are near completely removed as ground water moves
through the reactive media. Accompanying the reduction of sulfate concentrations is the accumulation in the solid-phase
of mackinawite (Fe1+xS) and microbial biomass with a phospholipid fatty extract signature consistent with the presence of
sulfate-reducing bacteria. Abiotic reduction of sulfate to sulfite, thiosulfate, elemental sulfur, or hydrogen sulfide are
known to be a sluggish processes at temperatures below 100°C (Trudingeretal., 1985). The slow kinetics of the abiotic
process is related to the eight-electron transfer required to reduce sulfate to sulfide.
In many subsurface systems, microbial reduction of sulfate is accompanied by oxidation of reduced organic carbon, e.g.,
SO/- + 2CH2O = 2HCO3- + HS" + H+ (5.13)
In Fe° barriers the amount of reduced organic carbon present would appear to be limiting and insufficient to account for
the amount of sulfate removed from solution and the quantity of iron sulfide precipitates observed in the solid phase.
Sulfate-reducing bacteria, such as Desulfovibrio desulfuricans, can also utilize hydrogen as a substrate for the reduction
of sulfate, e.g.,
SO42- + 4H2 + H+ = HS- + 4H2O (5.14)
In gate 2 of the DFC, we detected high concentrations of iron sulfide within the iron wall and comparatively low
concentrations of hydrogen. These observations are consistent with sulfate-reduction being an important sink for
hydrogen in Fe° barriers. A potential negative consequence of the presence of bacteria is biofouling. Proliferation of
bacteria could reduce the hydraulic conductivity of the barrier (Roh et al., 2000; Wilkin et al., 2003). On the other hand,
iron sulfide can be effective in enhancing the degradation of TCE and other halogenated aliphatic compounds (e.g.,
Butler and Hayes, 2000, 2001) and sulfur impurities in iron metal may also influence reactivity (e.g., Hassan, 2000). The
formation of iron sulfides in Fe° barriers, therefore, may be beneficial to their continued operation. In addition, compost-
based reactive barriers have been used to remove metal contaminants from ground water (Waybrant et al., 1998, 2002;
Benner et al., 1997). This treatment strategy is analogous to the use of anaerobic solid-substrate bioreactors for
removing metals from solution (e.g., Dvorak et al., 1992; Dairy, 1999). These systems rely on carbon-based media as
a substrate for sulfate reduction, sulfide production, and precipitation of insoluble metal sulfides. Sulfate-reduction could
also serve to increase the removal of metals in Fe° barriers through direct precipitation of metal sulfides, co-precipitation
with FeS, and/or adsorption onto iron sulfide surfaces.
Complete reduction of sulfate to bisulfide could lead to high levels of aqueous sulfide species in the absence of any
removal process. At near-neutral pH, iron monosulfides are insoluble so that all sulfide produced is likely to be
precipitated and removed from solution. At the DFC, we observed total dissolved sulfide concentrations of <1 ppm;
concentrations of sulfide in the solid-phase were as high as about 0.5 wt% S.
Field evidence suggests that microbial sulfate reduction is most effective in decreasing sulfate concentrations in low-flow
regimes. Iron barriers that experience higher flow-rates tend to show effective removal of sulfate (e.g., Morrison et
al., 2001; Kamolpornwijit et al., 2003).
Several studies have suggested that another possible sink for sulfate in Fe° barriers is the formation of sulfate green rust.
Geochemical modeling studies of iron corrosion described in Wilkin et al. (2002) suggest that the carbonate form of
96
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green rust is favored over the sulfate form at the Elizabeth City and DFC sites. X-ray diffraction results from these sites
are consistent with the formation of the carbonate form of green rust; sulfate green rust has not been detected. These
observations can be understood by considering the following exchange equilibrium based upon the anhydrous GR
components:
Fe6(OH)12S04 + CO/- = Fee(OH)12CO3 + SO/- (5.15)
so that,
K^5] = [SO/- / CO/-] / [Fee(OH)12C03 / Fee(OH)12SOJ (5.16)
Assuming ideal mixing relations in the solids and taking thermodynamic data from Bourne etal. (1999) for the anhydrous
carbonate and sulfate forms of green rust, we estimate ^515J = 103-1. Consequently, sulfate green rust (GR2) is expected
to be a primary iron corrosion product only when sulfate-ricn and bicarbonate-poor waters interact with zero-valent iron.
High-sulfate conditions are most likely to be present in impacted by the oxidative weathering of metal sulfide
deposits.
5.2.3
Field-deployed zero-valent iron systems effectively remove nitrate from ground water (e.g., Gillham et al., 1994;
McMahon et al., 1999; Liang et al., 2002). In this study, nitrate removal was consistently observed at the Elizabeth City
and Denver Federal Center PRBs over the initial five-year period of operation (Blowes et al., 1999b, Puls et al., 1999a,
Wilkin et al., 2002; this report). Laboratory investigations suggest that nitrate can inhibit the reduction of TCE and other
chlorinated ethenes through competition for reducing equivalents (e.g., Schlicker et al., 2000), however, as of yet, this
behavior has not been documented in the field.
At the Y-12 Pathway 2 PRB (Oak Ridge, TN), influent water contains nitrate at concentrations approaching 1,000 mg/L;
nitrate concentrations in effluent waters and in wells adjacent to the zero-valent iron zone are near or below analytical
detection limits (Liang et al., 2002). Unlike sulfate reduction that only proceeds via microbial respiration, reduction of
nitrate may proceed by either abiotic or biotic pathways (Siantaret al., 1996; Huang etal., 1998; Till etal., 1998; Gandhi
et al., 2002). Abiotic reduction is a pH-dependent process that results in the formation of ammonium and possibly nitrite
as an intermediate product. Rates of nitrate reduction in sterile systems are fast at pH < 4 (Huang et al., 1998), but
apparently reduction rates can be significant even at circumneutral pH (Siantar et al., 1996; Alowitz and Scherer, 2002,
but also Huang et al., 1998). The overall reaction may follow:
NO3- + 4Fe° + 10H+ = NH4+ + 4Fe2+ + 3H2O (5.17)
Huang et al. (1998) propose that electrons necessary for nitrate reduction are supplied directly from Fe° or indirectly
through an iron corrosion product (H2).
Data presented in Gandhi et al. (2002) suggest that the effective removal of nitrate observed in field PRB systems is
likely the result of significant biotic contributions. In biologically mediated systems, reduced organic carbon is frequently
the electron donorforthe redox transformation (Postma et al., 1991). In zero-valent systems, biotic reduction of nitrate
occurs by denitrifying bacteria that likely use cathodic H2 as an electron donor to respire NO3~ (e.g., Till et al., 1998;
Gandhi et al., 2002). Nitrogen gas is the principal product of biotic nitrate reduction rather than ammonia (Till et al.,
1998). A significant result of nitrate reduction from both abiotic and biotic mechanisms is increased iron corrosion,
leading to secondary mineral precipitation (Kamolpornwijit et al., 2003).
Ł2.4
Field evidence indicates that dissolved silica is removed by Fe° barriers (Gavaskar et al., 2002; Wilkin et al., 2003).
Ground-water concentrations of silica typically range between saturation values with respect to quartz and amorphous
silica (3 to 54 mg/L Si). Within iron walls, concentrations of Si typically fall below 1 mg/L. Using high-resolution
microscopy, Furukawa et al. (2002) observed silica predominantly associated with iron corrosion products. The role that
silica might play in passivated iron surfaces is not clear. However, recent long-term column studies suggest that silica,
carbonate, and natural organic matter co-solutes reduce the reactivity of Fe° (Vikesland et al., 2002).
Forms of SiO2 are not likely precipitating in iron walls because of their slow precipitation kinetics and because increasing
pH increases rather than decreases SiO2 solubility. One possibility is that silica is associated with magnesium in the clay
mineral sepiolite, Mg4(OH)2Si6O15-H2O. Sepiolite typically occurs as finely fibrous aggregates, but is less frequently
encountered in natural systems compared to the layered clay minerals. Sepiolite has a chain-like crystal structure of
continuous Si2O5 sheets with ribbons of Mg octahedra leaving channels that can incorporate water or organic molecules.
Figure 5.11 shows saturation indices of magnesium-bearing phases, sepiolite and brucite (Mg(OH)2), as a function of pH
for Elizabeth City ground water. With increasing pH, ground water approaches or clusters near the saturation points of
both sepiolite and brucite. Both phases, therefore, represent possible sinks for magnesium in iron walls (pH>9), and in
addition, sepiolite is a possible sink for silica. Appreciable buildup of these phases might only be expected in PRB
97
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2-
1-
0
-1-
-2-
-3-
-4-
-5-
-6-
-7-
-8-
-9-
-10-
-11-
-12-
undersaturated
• ML21 upgradient
• ML24 rridbarrier
A ML25 downgradient
PH
10
-4-
-6-
-8-
-10-
-12.
undersaturated
• ML21 upgradient
• ML24 rridbarrier
A ML25 downgradient
10
PH
Figure 5.11 Saturation indices of magnesium-bearing phases (brucite, Mg(OH)2; sepiolite, Mg (OH)2Si6O15-H2O) as a
function of pH in ground water from upgradient, iron wall, and downgradient sampling locations
(Elizabeth City PRB).
systems with greater influent magnesium concentrations than are encountered at the Elizabeth City or Denver Federal
Center sites.
5.2.5 Reactions with Oxygen
Concentrations of dissolved oxygen in influent ground water to Fe° barriers are rapidly consumed. Low concentrations
and the presence of dissolved ferrous iron complicate the accurate quantitation of dissolved oxygen in ground water
samples collected around iron walls. Oxygen is incorporated into a complex mixture of mineral oxides formed near the
surface of Fe° particles. The oxide film that develops on iron surfaces could be composed of magnetite, green rust,
maghemite (y-Fe2O3), hematite, goethite, lepidocrocite, or ferrihydrite. The oxide layer at the metal surface is expected
to evolve from a predominantly Fe(lll) phase where oxygen-containing solutions enter the Fe°zonetoa mixed-valent or
pure Fe(ll) phase under highly-reducing conditions expected in mid-barrier regions (e.g., Scherer et al., 1999;
Odziemkowski et al., 1998). High concentrations of dissolved oxygen entering Fe° barriers are especially problematic
due to rapid oxidation reactions, cementation by Fe(lll) materials, and plugging (e.g., Liang et al., 2000).
98
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5.3 Mineral Precipitation
Mineral precipitates in zero-valent iron PRBs can be classified by formation processes into three groups: 1) those that
result from changes in chemical conditions (i.e., change in pH, e.g., calcite); 2) those that are a consequence of microbial
activity (i.e., sulfate reduction, e.g., mackinawite); and 3) those that are the result of iron metal instability and corrosion
(e.g., magnetite). As noted in many publications, the formation of mineral precipitates in PRBs can impact system
performance through time. As minerals precipitate in iron walls they occupy volume and, therefore, reduce porosity and
permeability of the reactive zone. In this way, the hydraulic performance of PRBs (e.g., residence time, capture zone)
could degrade through time as the effective porosity of the iron wall approaches or exceeds that in the adjacent aquifer.
For example, preferential mineral accumulation in regions of a PRB resulting from higher inputs of dissolved solutes may
lead to increases in ground-water residence times. However, adjacent regions of the reactive barrier may experience
greater throughput and decreased residence times, potentially leading to contaminant breakthrough (Figure 5.12).
A second effect of mineral precipitation relates to the reactivity of materials placed in PRBs. As mineral precipitates
accumulate on iron surfaces (see Figures 3.24 and 4.24) it is expected that electron transfer processes, critical for the
degradation of chlorinated organic compounds, become less efficient. The formation of mineral precipitates has been
viewed generally as a process that limits the long-term performance of reactive barriers for ground-water cleanup. Yet
some corrosion products that deposit on the surfaces of iron particles may also contribute to the overall treatment
effectiveness of reactive barriers (e.g., Butler and Hayes, 1998, 1999, 2000; Lee and Batchelor, 2002; Furukawa et al.,
2002). For example, iron sulfides, magnetite, and green rust minerals can chemically transform chlorinated organic
compounds.
5.3.1 Pore Volume Reduction
The infilling of pore space by mineral precipitates can be assessed by theoretical modeling efforts and through direct field
measurements. The development of models of mineral precipitation in PRBs is of great interest because these models
can be used as predictive tools during remedial investigations at contaminated sites (e.g., Blowes and Mayer, 1999;
Mayer et al., 2001; Yabusaki et al., 2001; Liang et al., 2003). Field measurements are critical in verifying model
predictions at specific sites. Based upon the results of this and other studies, volume loss in zero-valent iron systems
results primarily from the formation of mineral precipitates containing carbon, sulfur, and iron.
pei, Ki, Gradient T
Seepage velocity
PRB
Decreased T
Zone of increased
buildup, increased T
Decreased T
Figure 5.12 Conceptual model of the impact of mineral and biomass accumulation to PRB hydraulic performance
(p =effective porosity; K=hydraulic conductivity; i=residencetime).
99
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After five years of mineral accumulation, concentrations of inorganic carbon in core materials from the Elizabeth City and
Denver Federal Center PRBs range from <1 to about 14,100 u,g/g. This concentration range can be directly related to
a fractional porosity reduction, i.e., the reduction in the fraction of the total volume occupied by pore space. In
Figure 5.13, fractional porosity reduction resulting from formation of carbonate minerals is plotted as a function of the
concentration of inorganic carbon. An inorganic carbon concentration of 15,000 u,g/g would result in a decrease of the
fractional porosity by 0.13 to >0.5, depending on how the carbon is distributed in the solid phase (i.e., siderite or
carbonate green rust, respectively). The least amount of pore infilling occurs when carbon is partitioned into minerals
with low molar volumes, such as siderite, aragonite, and calcite (Table 5.2). Greater porosity loss is a consequence of
the formation of carbonate green rust and iron hydroxy carbonate, materials with comparatively high molar volumes
(Figure 5.13; Table 5.2). After the first five years of operation at Elizabeth City, the maximum loss of fractional pore
space near the upgradient iron/aquifer interface due to the formation of inorganic carbon precipitates is estimated to be
about 0.07. At the Denver Federal Center, the maximum loss of fractional pore space due to the formation of inorganic
carbon precipitates is estimated to be about 0.17 after the first five years of operation.
The main sulfur-bearing mineral in zero-valent iron PRBs is mackinawite (FeS) as seen in this study and at other sites
(e.g., Phillips et al., 2000; Roh et al., 2000). The low molar volume of mackinawite results in little loss of porosity (<0.05)
even at the highest concentrations of total sulfur observed in this study (7,520 u,g/g; Figure 5.14). If through time
mackinawite were to completely transform to pyrite, the loss of pore space would be even lower due to the low molar
volume of pyrite (Table 5.2). Sulfur partitioned into sulfate green rust would result in much more significant porosity
reductions (Figure 5.14).
Elizabeth City PRB after 5 y
Denver Federal Center PRB after 5 y
0.5.
10000 20000 30000
Measured inorganic carbon, jig/g
40000
0.4-
0.3-
0.2-
0.1-
0.0-
Aragonite
Calcite
Siderite
Carbonate green rust
Iron carbonate hydroxide
10000 20000 30000
Inorganic Carbon, |ig/g
40000
Figure 5.13 Fractional porosity reduction as a function of inorganic carbon concentration in the solid phase. The
lines represent volume loss due to the accumulation of a carbon-bearing phase assuming that all carbon
is present in that phase. Fractional porosity reduction is calculated assuming an initial porosity of 0.50
and iron density of 7.0 g/cm3. The range of observed inorganic carbon concentrations after 5 years at
the Elizabeth City and the Denver Federal Center PRBs are shown.
100
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Table 5.2. Molar Volume and Density of Mineral Precipitates
Phase
Calcite
Aragonite
Siderite
Green Rust -
C03
Iron Hydroxy
Carbonate
Mackinawite
Pyrite
Green Rust -
SO4
Iron
Magnetite
Hematite
Goethite
Ferrihydrite
Formula
CaCO3
CaCO3
FeCO3
Fe6(OH)12C03
Fe2(OH)2CO3
FeS
FeS2
Fe6(OH)12S04
Fe
Fe3O4
Fe2O3
FeOOH
Fe(OH)3
Molecular
Weight
(g/mole)
100.09
100.09
115.86
599.17
205.72
87.91
119.98
635.23
55.85
231.54
159.69
88.85
106.87
Density
(g/cm3)
2.71a
2.95 a
3.96a
3.5°
3.65 c
4.1°
5.02 a
3.5°
7.0a
5.18a
5.26a
4.37a
3.1°
Molar Volume
(cm3/mole)
36.93
33.93
29.26
171.19
56.36
21.44
23.90
181.49
7.98
44.70
30.36
20.33
34.47
Molar volume = MW/density
a Hurlbut and Klein (1977); estimated;c Erdos and Altorfer (1976);d Vaughan and
Craig (1978)
Corrosion reactions in which iron reacts to form iron oxides such as magnetite, hematite, goethite, or ferrihydrite also
result in volume increases and porosity reductions. For example, the molar volume change of the reaction (A V) from Fe°
to magnetite can be computed from the reaction:
Fe° + 4/3HO = 1/3 Fe3O, + 4/3 H,(g)
(5.18)
by using the molar volumes (Vm) of Fe° and magnetite listed in Table 5.2. The molar volume change of reaction is given
by:
AVr = 1/3(Vm magnetite) - (Vm Fe°) = 6.9 cm3/mol
(5.19)
Given the assumption that iron is conserved in the solid phase, transformation reactions of iron metal to iron hydroxides,
oxyhydroxides, and oxides all have positive molar volume changes. The fractional porosity reduction associated with
various iron transformations are plotted in Figure 5.15 as a function of extent of reaction or extent of transformation.
Although it is well documented that magnetite, for example, is a product formed in iron walls, the overall extent of the
transformation has not been directly determined or estimated. Based upon microscopic examination, iron particles
collected from the Elizabeth City and DFC PRBs (after five years) are dominantly composed of Fe° and not iron oxidation
products. An average rate of iron metal corrosion in ground water is estimated to be 0.6 mmol/kg d (Reardon, 1995). At
this rate of corrosion, after five years the extent of reaction to magnetite is estimated to be 0.02. If ferrihydrite is the end
product instead of magnetite, the extent of reaction is estimated to be 0.06. Consequently, the extent of reaction after
5 years is likely to be less than 0.3 and probably much less than 0.1. As shown in Figure 5.15, the volume loss
accompanying the transformation to magnetite is not expected to exceed 0.1 after five years. It should be noted that
where rapid oxidation occurs due to the influx of ground water with high levels of dissolved oxygen (>2 mg/L), reaction
to form goethite or ferrihydrite has been documented. Figure 5.15 shows that even partial conversion of Fe° to
ferrihydrite can dramatically impact the porosity of zero-valent iron systems.
101
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Elizabeth City PRB after 5 y
Denver Federal Center PRB after 5 y
10000 20000 30000
Measured sulfur, |j,g/g
40000
50000
Mackinawite
Sulfate green rust
Pyrite
10000
20000 30000
Sulfur, ng/g
40000
50000
Figure 5.14 Fractional porosity reduction as a function of sulfur concentration in the solid phase. The lines represent
volume loss due to the accumulation of a sulfur-bearing phase assuming that all sulfur is present in that
phase. Fractional porosity reduction is calculated assuming an initial porosity of 0.50 and iron density of
7.0 g/cm3. The range of observed sulfur concentrations at the Elizabeth City and the Denver Federal
Center PRBs are shown.
Fe-to-Magnetite
Fe-to-Hematite
Fe-to-Goethite
Fe-to-Ferrihydrite
Figure 5.15
0.4 0.6
Extent of Reaction
Fractional porosity reduction as a function of the positive molar volume change as iron metal reacts to
form magnetite, hematite, goethite, and ferrihydrite.
102
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Ł3.2 of
The previous calculations examining porosity reductions in iron walls as a consequence of mineral precipitation relate
primarily to potential impacts to the hydraulic performance of PRBs. However, these calculations provide no direct
insight about how mineral precipitation affects or could interfere with contaminant removal processes. In order to
examine changes in the reactivity of zero-valent iron as a function of exposure time, we conducted batch experiments
using fresh iron (Peerless Iron) and materials that were collected from the Elizabeth City and Denver Federal Center
PRBs after five years of exposure to ground water. In the batch tests, iron samples used were collected from upgradient
and midbarrier locations as summarized in Table 5.3. The water composition used in the batch tests was prepared using
reagent grade sodium bicarbonate, calcium sulfate, magnesium sulfate, magnesium chloride, potassium chloride, and
hydrochloric acid to approximately match ground-water compositions encountered at the two PRB sites {Table 5.4).
Solutions were deoxygenated by purging with high-purity nitrogen gas for one hour.
Stock solutions were prepared using high purity trichloroethylene (99+%, Aldrich), 1,1,1-trichloroethane (anhydrous,
99+%, Aldrich), and 1,1-dichloroethene (99%, Aldrich). Experiments to determine the kinetics of VOC degradation and
Cr removal were carried out in 45-mL glass VOA vials, each containing 10 g of iron and filled with freshly prepared
synthetic ground water. The VOA vials were without headspace with Teflon-lined screwcaps. Next the reaction
vessels were injected with a volume of the VOC stock solution and placed in a rotary shaker at 100 rpm at room
temperature (23 ± 1 °C). Visual observations suggest that the gentle mixing did not result in any physical abrasion of the
iron particles or corrosion products. Periodically vials were withdrawn for sample collection and measurement of pH,
oxidation-reduction potential, sulfate, chloride, calcium, magnesium, iron, sodium, potassium, chromium, and VOCs.
Control experiments (without iron) showed no loss of organohalides or metals over the relevant experimental timescales.
Results of the batch experiments show that field-exposed iron samples from both midbarrier and upgradient locations are
able to remove contaminants from solution at rates comparable or, even better, to those observed in systems containing
unreacted zero-valent iron (Figure 5.16; Table 5.5). The initial removal rate of chromium is actually greater for field-
Table 5.3. Samples Used in Batch Reactivity Tests
Sample
Unreacted
Peerless Iron
EC050801-3-1
EC050901-9-1
C1 -1-71 701 -19
C2-3-71801-4
Location
-8/+50 mesh
size
Elizabeth City,
upgradient
Elizabeth City,
midbarrier
Denver Federal
Center, Gate 1 ,
midbarrier
Denver Federal
Center, Gate 2,
upgradient
Inorganic Sulfur, (ag/g
Carbon, i_ig/g
<15 <5
4633 3880
100 100
87 318
2100 1785
PLFA, pm/g
ND
ND
ND
ND
455
103
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Table 5.4. Composition of Water Used in Batch Reactivity Tests
Component Synthetic Elizabeth City Synthetic Denver Federal
Ground Water (mg/L) Center Ground Water (mg/L)
Ca
Mg
Na
K
Cl
SO4
HCO3
pH
10
7
31
4
17
30
82
6.0
110
32
163
4
68
300
433
7.2
exposed iron compared to unreacted iron {Figure 5.16). The added component of mineral precipitates (carbonates and
sulfides) and microbial biomass to zero-valent iron apparently increases the uptake rate of chromium, probably due to an
increase in the number of available sorption sites. Increased chromium removal was also observed in experiments with
field-exposed iron materials from the DFC (where chromium is not a ground-water contaminant).
Table 5.5 presents a summary of rate data for reactions of TCE and 1,1,1 -TCA. Values of log kSA in Table 5.5 are based
on pseudo-first-order reaction kinetics. In general, log #CSA values for TCE are independent of whether the iron used was
fresh or contained quantities of mineral precipitates and microbial biomass. The linearity of ln[TCE] vs. time plots
decreases in the order unreacted iron>midbarrier iron>upgradient iron, as indicated by the standard regression
coefficient. This trend suggests that removal processes are more complicated in materials containing authigenic
components. Results for 1,1,1-TCA give log feSA values that range from-3.1 to-4.0. The effect that mineral precipitates
play is still unclear as the slowest and fastest kinetics were observed, respectively, in systems containing upgradient iron
from the DFC and Elizabeth City PRBs.
The batch experiments indicate that zero-valent iron retains reactivity and the ability to remove chlorinated ethenes,
ethanes, and hexavalent chromium even after long-term exposure times to ground water. Interestingly, the ability of
zero-valent iron to remove chromium from ground water appears to improve with time. The Elizabeth City PRB is
expected to remain effective for chromium removal for another five-year period at a minimum. Observations of
contaminant breakthrough in field PRBs may be more directly tied to decreases in hydraulic performance (i.e., system
residence time, plume bypass) rather than loss of reactivity of zero-valent iron.
5.4
The oxidation or corrosion of zero-valent iron may be stimulated or inhibited by microorganisms. From a subsurface
ecological perspective, metallic iron represents a significant energy reservoir. Due to the limited solubility of oxygen in
ground water and the rapid reduction of molecular oxygen by Fe(0) and Fe(ll), PRBs usually exist as anaerobic
environments. Under anaerobic conditions molecular oxygen-driven chemical corrosion rates may be reduced, but
biologically mediated anaerobic corrosion may occur at rates exceeding those seen under oxygenated conditions. In the
absence of oxygen, protons may serve as electron acceptors and allow for the formation of oxidized iron species such
as Fe(ll).
The enhancement of anaerobic corrosion and the formation of dissolved Fe(ll) and hydrogen is not necessarily
detrimental to PRB performance. If the target contaminant such as Cr(VI) is reduced by Fe(ll) as well as Fe(0),
production of aqueous Fe(ll) could increase the of the reaction/treatment zone as compared to the surface contact
of Fe(0) alone. However, the utilization of dissolved hydrogen may result in bacterial growth and biofilm formation.
The development of this biofilm in a PRB may be detrimental to performance through several mechanisms. Biofilm
growth in a porous medium may reduce the total volume and the average size of the pores (e.g., Taylor et al., 1990a;
Thullner et al., 2002). Changes in PRB hydraulic conductivity, the masking of active sites, the removal of active chemical
species, mineral precipitation, production of gas bubbles, and the competition for reducing equivalents are processes
mediated by bacteria that could negatively affect PRB performance. Conversely some microbial processes could
enhance PRB performance. In some instances, bacteria may be more effective at contaminant transformation or may
degrade compounds unaffected by PRBs. It is, therefore, evident that a clear understanding is needed of microbial/PRB
interactions for the design and efficient operation of PRBs.
104
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3500-
3000-
2500-
=d 2000-
-------
Table 5.5. Summary of Rate Data for Reactions of TCE and 1 ,1 ,1-TCA with Zero-valent Iron (Unreacted and
Collected from Field PRBs)
System TCE 1,1 ,1-TCA
_ log frSA(L/h.m2) ff2 log ^(L/h-m2) ff2 _ n
Elizabeth City
(GW) -3.77 0.885 -3.86 0.949 10
Unreacted
Peerless Fe
Elizabeth City
(GW) -3.67 0.841 -3.66 0.935 7
EC Iron
midbarrier
Elizabeth City
(GW) -3.72 0.781 -3.12 0.978 6
EC Iron
upgradient
DFC (GW)
Unreacted -3.83 0.988 -3.52 0.998 7
Peerless Fe
DFC (GW)
DFC Iron -3.81 0.962 -3.82 0.990 9
midbarrier
DFC (GW)
DFC Iron -3.93 0.906 -4.06 0.884 8
upgradient
Midbarrier and downgradient regions of the reactive barrier are, in most cases, free of microbial biomass. PLFA analysis
suggests that a diverse assemblage of microorganisms colonize zero-valent iron systems at Elizabeth City and the
Denver Federal Center. The formation of comparatively high-density, contaminant-reactive iron sulfides is one indirect
consequence of microbiological activity (sulfate reducing bacteria).
The visual appearance and results of PLFA analyses from iron core materials collected from DFC gate 2 suggests that
high levels of microbial biomass could significantly reduce permeability of PRBs. Previous laboratory investigations have
documented this effect (e.g., Taylor and Jaffe, 1990b; Vandevivere and Baveye, 1992). Low-flow systems may be
problematic in this regard. Although it is clear that microorganisms colonize zero-valent iron systems, an outstanding
question remains as to whether microorganisms play a direct role in reducing contaminant concentrations. Additional
research is needed to address this question.
A comparison of the average PLFA distribution among samples within reactive iron materials collected from Elizabeth
City, DFC gate 1, DFC gate 2, and the Moffett Field PRB (data from Gavaskar et al., is shown in Figure 5.17. PLFA
distributions in these four PRBs are broadly comparable. The Elizabeth City PRB is comparatively enriched in normal
saturated (Nsats) structural groups that are ubiquitous in both prokaryotic and eukaryotic organisms. The broad
similarity suggests that similar microbial populations colonize zero-valent iron systems. PLFA profiles in the reactive
media tend to mirror profiles found in the aquifer materials sampled immediately upgradient to the reactive media;
however, biomass concentrations are significantly greater within the iron zones. Indigenous microbial communities
(especially anaerobes) appear to be stimulated by the placement of iron in the subsurface.
106
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Microbial Biomass - PLFA Distribution
LL
Q
N=18
<3500 pm/g
CM
"5J
O
O
u_
D
N=44
<4920 pm/g
ui
=40
<2600 pm/g
N=2
<240 pm/g
n Gram+/ anaerobic
Gram-
• Gram- (Monos)
D Anaerobic metal
reducers
DSRB/
Actinomycetes
• Genera Nsats
From Gavaskaretal., 2002
Figure 5.17 Comparison of PLFA distribution in four iron walls.
5.5 Hydrogeological Issues
The majority of recent research concerning PRBs has been directed at understanding geochemical processes involving
different reactive materials as well as issues relating to predicting long-term performance. In comparison, fewer studies
have explored issues of ground-water hydrology and flow behavior in PRBs. Clearly hydraulic and geochemical
processes in PRBs are interrelated (see Tratnyek et al., 1997; Yabusaki et al., 2001; Mayer et al., 2001; Das, 2002). An
understanding of the hydrologic properties of aquifers and reactive media, such as porosity, hydraulic conductivity
distribution, hydraulic gradients, and flow velocities, heavily factor into the design and ultimately the success of PRB
installations (Gavaskaretal., 1998). Two critical issues in system design that require an understanding of the spatial and
temporal aspects of site hydrology are: 1) plume capture, i.e., ensuring that contaminated ground-water plumes are
directed through the reactive material without bypass, underflow, or overflow; and, 2) residence time, i.e., ensuring that
ground water remains in contact with the reactive media for an adequate period of time to allow for the removal or
transformation of contaminant compounds.
Aquifer heterogeneity in chemical and physical properties is typically present to some degree at all sites. Such
heterogeneities will result in variable contaminant flux across the influent area of the PRB and variable residence time
requirements for contaminant treatment. Reliance on the use of bulk or averaged geochemical and hydraulic parameters
may potentially result in inadequate system designs (Gavaskar et al., 2002). Several studies provide some insight into
how aquifer heterogeneity can impact the performance of PRBs. Eykholt et al. (1999) concluded that seepage velocities
within a homogeneous reactive media are principally controlled by heterogeneity in the downgradient aquifer. They
showed that variability in hydraulic conductivity spanning two orders of magnitude resulted in variability in flow velocities
and residence times within the barrier that span one order of magnitude. Studies by Tri-Agency partners (Kamolpornwijit
et al., 2003; Moline et al., 2002) show that treatment of high nitrate and TDS ground water has lead to increased mineral
precipitation, which facilitated the development of heterogeneous flow, in addition to the initial heterogeneity present.
This additional preferential flow over a year of PRB operation could cause ground-water bypass. In another modeling
study, Benner et al. (2001) showed that localized or narrow high conductivity zones within the aquifer lead to greater
preferential flow within the reactive media (see also Gupta and Fox, 1999). Benner et al. (2001) suggest that less
variable flow will be attained using thicker, homogeneous barriers.
One of the most critical issues that must be addressed in the design of PRBs is the selection of the appropriate barrier
width. The barrier width must provide sufficient contact time to insure that contaminants are degraded to target levels.
107
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The barrier width must be selected based upon the range of ground-water flow velocities expected and the desired extent
of contaminant removal for a given reactive material. Tratnyek et al. (1997) used laboratory derived kinetic data
pertaining to the degradation of various chlorinated halogens by zero-valent iron and a one-dimensional transport model
to estimate the minimum barrier width required for specified contaminant reductions as a function of ground-water flow
velocity. This type of analysis is critical for providing a quantitative basis for system design to accompany laboratory
feasibility testing.
Gavaskar et al. (1998) discuss the tools and methods that may be used to maximize the probability of success in
hydraulic aspects of PRB performance. They emphasize the importance of thorough site characterization and ground-
water flow modeling. A principal goal of site characterization is the development of a detailed understanding of site
geology, hydrogeology, and contaminant distributions. Site characterization efforts should be to the level of capturing
seasonal and spatial variability. Ground-water modeling efforts incorporate site characterization data to simulate water
and contaminant transport. Modeling results will directly feed into a system design that ensures plume capture and
adequate residence time for contaminant removal.
Hydraulic aspects of performance monitoring generally include water level surveys, hydraulic conductivity measure-
ments, and determinations of seepage velocity. Water level surveys provide information on ground-water gradients and
capture zones of PRBs. Water level surveys were regularly conducted at the Elizabeth City site on a quarterly basis.
The results of water level surveys from June 1997, 1998, and 1999 are shown in Figure 5.18. Inspection of the water
level map shows that the primary flow direction is across the PRB. In general, the hydraulic gradient in June varies from
about 0.001 to 0.004, and this range further captures annual variability in the hydraulic gradient at this site. Based on
hydraulic conductivities measured from slug tests and the hydraulic gradient obtained from water level measurements,
a typical ground-water velocity of 0.5 ft/day and a typical residence time of four days are estimated.
McMahon et al. (1999) discussed the hydraulic performance of the Denver Federal Center PRB. Installation of the
funnel-and-gate system at the DFC resulted in the mounding of ground water on the upgradient side of the sheet pile due
to insufficient flow through the system. The buildup of ground-water levels on the upgradient side of the PRB raised
concerns about the increased potential for ground-water bypass; flow under, over, or around the PRB. Follow-up studies
suggest that underflow and overflow are not occurring, but that some bypass occurs around the southern side of gate 1
(Pacific Western Technologies, 2000). A downgradient distribution ditch was installed in late 1998, connecting gates 1
and 2, to decrease the ground-water mound. Although this attempt was unsuccessful in decreasing the head difference
across the gates, the trench did greatly decrease the water levels within gates 1 and 2 (Pacific Western Technologies,
2000). Before and after installation of the distribution trench, the head differential across gate 2 has averaged
approximately 7 ft. The lowering of water level in gates 1 and 2 suggested that only partial hydraulic connectivity existed
at the upgradient aquifer/iron interface. Further studies by FHWA and the GSA were initiated to understand the hydraulic
and performance issues at the DFC, including parts of the investigation described in this report.
It is believed that the head differential in gate 2 is a result of low permeability zone that was produced by backfilling pre-
excavation trenches with muddy material. Pre-excavation was required in order to install the sheet pile. Alternatively, a
smear zone of fine materials could have resulted as a consequence of the installation and removal of sheet piling
installed for gate construction. In either case, flow velocity through gate 2 is reduced compared to that through gate 1.
Characterization studies described in this report suggest that flow through gate 2 is reduced over a depth interval from
about 5.5 to 8 m below ground surface. This reduced flow zone may be related to the low permeability zone or smear
zone that developed during construction of the gate.
108
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Pasquotank River
Iron Wall
1997
Former Chroma Pl
ating Shop |
(contour interval = 0.05 ft)
Pasquotank River
Iron Wall
MW18
June 1998
(contour interval = 0.05 ft)
I Former Chrome Plating Shop I
Pasquotank River
Iron Wall
June 1999
I Former Chrome Plating Shop I
(contour interval = 0.05 ft)
Figure 5.18 Water levels in Elizabeth City monitoring wells.
109
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110
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6.0 of
6.1 An for &
of
The use of permeable reactive barriers for the restoration of contaminated ground water has evolved from innovative to
accepted, standard practice, for the containment and treatment of a variety of contaminants in ground water. Like any
remedial technology, the decision to use PRBs will be conditioned by the nature of the natural system, the target
contaminants, and the treatment objectives. In the past 10 years, more than 100 sites have implemented this technology
to treat chlorinated solvent compounds, fuel hydrocarbons, and various inorganic contaminants in ground water. As with
any technology used to treat or extract contaminants in the subsurface, successful implementation will be contingent on
effective site characterization, design, and construction. Our studies on long-term performance of the technology at a
number of sites in the United States have shown the following with respect to ensuring (designing) and verifying
(monitoring) that the PRB meets performance objectives:
• Adequate site characterization is necessary on the scale of the PRB. Site characterization approaches,
typical of remedial feasibility investigations, are oftentimes not adequate. Additional localized characteriza-
tion of the plume distribution in four dimensions (including time), understanding of local hydrogeology, and
knowledge of the geochemistry of the site is required.
• Understanding of site hydrology has emerged as the most important factor for successful implementation.
This is not surprising given the nature of the technology. The PRB must be located to intercept the plume.
Once located in the subsurface, it cannot be moved, so an understanding of how the PRB will impact the
prevailing flow patterns is important. It is imperative that the selected design allow for capture of the plume
in its present configuration, as well as allow for variations in flow direction, depth, velocity, and concentra-
tions of contaminants, which may vary over time.
• There is a need to develop contingency plans in case a system fails to meet design objectives. This requires
specification of design criteria, performance objectives, and what constitutes a failure in order to clearly
trigger the activation of contingency plans, i.e., alternative technologies or remedies to the installed PRB
system.
Performance goals should target the adequacy of plume capture and contaminant treatment such that acceptable
downgradient water quality is achieved in a reasonable time frame. Short-term objectives generally involve establish-
ment of adequate residence time of the contaminant(s) in the reactive media to achieve treatment goals, while long-term
objectives revolve around longevity or lifetime expectations for the system, which in turn, affect cost.
Specific criteria need to be established for all these concerns, and both parties (site owner and regulator) need to be
clear on what triggers contingency plans. Maximum contaminant level (MCL) concentrations for contaminants in ground
water are often used as criteria at points of compliance. This approach becomes complicated when contaminant levels
already exceed goals at the point of compliance, and meeting these goals is contingent on desorption of residual
contaminants or "flushing" over time. A time period needs to be specified in such a case that is reasonable given site
characteristics and known contaminant behavior.
Performance goals are usually developed for the site as a whole and contingent upon plume and barrier location relative
to compliance points and/or site boundaries. The performance goals can be numeric, regulatory-driven targets and may
have system design features including remedial measures, such as natural attenuation, downgradient of the barrier
location. The time horizon for sustaining performance goals will depend on site-specific factors related to chemical,
physical and microbiological processes, but also such factors as the extent of source removal, source containment or
expected lifetime of the source as a plume generator.
In many cases, contingency plans are required in the event that the PRB fails to meet performance criteria. Such plans
may range from minor modifications of the PRB to use of an alternative technology. If the PRB fails to capture a portion
of the plume, an extension to the PRB may be prescribed. If concentrations of contaminants exiting the PRB are higher
111
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than expected, then an additional wall may be required downgradient of the first wall. In some installations, monitored
natural attenuation of contaminants downgradient of the wall is expected and designed into the system to help meet
compliance goals,
6.2
As with any ground-water remediation technology, adequate hydrogeologic characterization must be done to understand
flow patterns and the distribution of the contaminant plume. This is particularly important for PRBs as the treatment
system is immovable or passive, yet it must intercept and capture the contaminant plume for effective treatment.
Information to be obtained includes advective velocity parameters such as the gradient, the hydraulic conductivity,
porosity, and other parameters collected as part of a hydrogeologic characterization program. It is also important to
understand temporal changes in flow direction and flux due to processes such as recharge, pumping of adjacent wells or
other disturbances. Observed changes in flow direction at the Elizabeth City site, for example, have ranged as high as
25 degrees from time to time.
In addition to the hydrology, the stratigraphy and lithology of the site is important to understand and will dictate effective
PRB design. If a low permeability layer exists at the site, the PRB can be keyed into this layer. If it does not exist, then
a "hanging wall" design must be chosen which may add to the uncertainty of plume capture. If the site has low
permeability layers through which the PRB must be constructed, care must be taken during construction to avoid
smearing of such layers. This could impact hydraulic contact between the formation and the reactive media. A thorough
understanding of site stratigraphy Is especially important or helpful In choosing a particular construction method. Use of
sheet piling to construct a reactive "gate" may not be a good choice where low permeability layers exist because of the
smearing potential and increased difficulty in reestablishing good hydraulic contact between aquifer sediments and the
reactive zone.
Characterization of contaminant concentrations In four dimensions is required for successful implementation of a PRB.
In addition to knowledge of the plume in the three-dimensional space, it is also Imperative to understand variability in
plume shape and direction over time. Plumes deviate In direction and location over time and may change shape due to
attenuation, degradation, mixing with other plumes, dilution, recharge, and other natural and anthropogenic-Induced
disturbances.
PRBs are often located within plumes. This requires some understanding of the impact of construction on plume
behavior, both upgradient and downgradient of the barrier. For example, it is essential to verify that hydraulic contact
between the plume and reactive media is established. If a PRB is located below an impermeable surface structure such
as a parking lot, will the surface be repaved immediately or will recharge be allowed to occur over the PRB? Some
understanding of natural attenuation processes at the site is Important in being able to interpret the subsequent response
of the natural system to the presence of the PRB. This is most often manifested in trying to estimate how long the
downgradient aquifer will require to achieve cleanup goals. A key question to address is the length of time required
before contaminants located downgradient of the barrier flush out of the sediments or degrade naturally,
6.3 Of
Geochemical characterization of sites for PRBs Is Important for optimizing the design and performance of a PRB and for
predicting longevity. The lifetime of a PRB will depend on the hydrogeochemlcal nature of the site, flow rate, and
contaminant flux, among other factors. It Is known that high carbonate waters, high nitrate waters, high dissolved organic
carbon waters, or waters with generally high total dissolved solids will have shorter life expectancies than what might be
considered for "typical" or "average" composition waters. Decreased life expectancy may be caused by competition for
reaction sites, loss of reactive sites due to rapid corrosion or fouling, or precipitation of inorganic minerals due to changes
in geochemistry caused by the presence of the reactive media with subsequent loss of permeability.
If zero-valent iron is the reactive media, corrosion reactions together with mineral precipitation will eventually result in
loss of permeability and/or reactivity resulting in decreased performance to the point where performance goals are no
longer met. Long-term performance studies have documented decreased reactivity at some sites over time as well as
loss of porosity, which can affect residence time of contaminants in the reactive media. For inorganic contaminants,
such as chromium and arsenic, removal capacities have been calculated for zero valent iron PRBs. These capacities
can be quickly reached If waters are rich In species that compete for reaction sites.
Monitoring of geochemlcal parameters (e.g., pH, Eh, dissolved oxygen [DO], specific conductance, terminal electron
acceptors) In iron-based PRBs will verify that the system functions as designed once Installed. Within the zero valent
iron barriers, the change In these parameters Is strikingly different than the natural system (e.g. Eh ~ -400 mV, pH > 9,
DO = 0). Trends In these parameters may signal changes in system performance, but no clear correlations between
these parameters and decreased performance have been observed to date. Long-term trends for these parameters are
consistent with contaminant trends observed in both sites studied and reported on in this document. Spatial and
temporal variations in the concentration distribution of terminal electron accepting species (e.g., sulfate), specific
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conductance, and Eh suggest that both anaerobic iron corrosion and microbial activity play important roles in controlling
the oxidation-reduction potential in iron barriers. Low Eh values (<-100 mV relative to the SHE) and decreases in the
specific conductance of ground water between upgradient contaminant plumes and sampling points within reactive iron
media are consistently indicative of normally operating PRB systems. Anomalous behavior in these parameters may be
useful indicators of problems associated with construction, as observed in gate 2 at the Denver Federal Center.
Performance monitoring also focuses on the rate of mineral buildup within the reactive media. This may lead to
decreased permeability and clogging. For zero valent iron systems, the reactive media is a long-term sink for C, S, Ca,
Si, Mg, and N. The buildup of mineral precipitates is related to influent ground-water chemistry and flow rate. Mineral
precipitates and microbial biomass accumulate the fastest near the upgradient aquifer-Fe° interface. Porosity loss in the
iron media due to precipitation of inorganic carbon and sulfur minerals can be estimated by integrating the concentrations
of inorganic carbon and sulfur as a function of distance in the iron and estimating the volume loss by using the molar
volumes of zero-valent iron, calcium carbonate, iron carbonate, and iron sulfide. The rate of mineral accumulation and
the rate of iron corrosion varies spatially, therefore, so does the rate of porosity infilling. Porosity loss estimates have
ranged from 1 to 4 % per year in this study. Based on these estimates, the average porosity of the PRB at Elizabeth City
would not be expected to approach that of the surrounding aquifer for 15 to 30 years. The highest concentrations of
mineral precipitates and rates of porosity loss are found adjacent to upgradient interfaces, and at Elizabeth City, there is
a vertically localized zone which corresponds to the higher specific conductance portion of the plume. As these zones
lose porosity, flow may be diverted to other locations along the face of the wall, thus extending lifetime estimates based
on worst-case scenarios. As corrosion minerals form on the surface of the iron media, reactive surfaces are coated,
presumably decreasing the effective reactive surface area. However, corrosion products formed include some minerals
which themselves are highly reactive and capable of transforming inorganic and organic contaminants into immobile or
non-toxic species. This phenomenon must also be factored into lifetime projections.
Microbiological impacts are important to understand in order to better predict how long PRB systems will remain
effective. The presence of a large reservoir of iron, coupled with abundant substrate supports the metabolic activity of
iron-reducing, sulfate-reducing, and/or methanogenic bacteria. These populations may have either beneficial or
detrimental effects on system performance. Enhanced biodegradation of contaminants is possible where growth is
stimulated by the presence of the reactive media, but biofouling may lead to permeability reduction within the reactive
media or immediately upgradient. Additional research is required to assess changes in microbial ecology associated with
the installation of these systems. Indications from Elizabeth City are that enhanced treatment some distance away from
the PRB occurs, and it will be interesting to follow this development over extended periods of time.
While long-term performance observations of the Elizabeth City and Denver Federal Center site are now approaching
seven years, there has still not been sufficient time to adequately predict the lifetime of these PRBs. It is clear that
lifetimes exceeding 10 years are reasonable to expect, and they may function adequately for much longer. Continued
studies are needed to better predict longevity based on ground-water composition, flow rate, and contaminant flux.
6.3.1 for
The research conducted to date on the long-term performance of permeable reactive barriers has uncovered a number
of new questions that require further investigation. In addition, several complex outstanding issues still require further
study. Topical areas of research are outlined below where continued efforts will lead to more successful implementation
of PRB technology for ground-water remediation.
« Tools are needed to more quantitatively relate ground-water chemistry to porosity loss, changes in hydraulic
performance, and changes in Fe° reactivity.
• An improved understanding is needed of how the reactivity and removal capacity of iron-based media
changes with time and continued ground-water exposure, with respect to various inorganic and organic
contaminants of concern, and over a range of hydrogeological and geochemical conditions.
« Research is needed on the microbial ecology of iron-based PRBs, especially with respect to potential
benefits and deleterious effects, and the factors that control biomass accumulation.
« A better understanding and development of approaches are needed to determine flow patterns through
reactive barriers in three-dimensions.
• Application of the PRB technology in conjunction with other subsurface treatment technologies is needed
(e.g., monitored natural attenuation, source zone treatment).
« New methods are needed for regeneration of the reactive medium in situ, without having to excavate and
entirely or partially replace the PRB.
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7.0
Agrawal, A, and Tratnyek, P. G. (1996). Reduction of nitro-aromatic compounds by zero-valent iron metal. Environmental
Science and Technology, v. 30, p. 153-160.
Agrawal, A., Ferguson, W. J., Gardner, B. O., Christ, J. A., Bandstra, J. Z., and Tratnyek, P. G. (2002). Effects of
carbonate species on the kinetics of dechlorination of 1,1,1-trichIoroethane by zero-valent iron. Environmental
Science and Technology, v. 36, p. 4326-4333.
Alowitz, M. J. and Scherer, M. M. (2002). Kinetics of nitrate, nitrite, and Cr(VI) reduction by iron metal. Environmental
Science and Technology, v. 36, p. 299-306.
Arnold, W. A. and Roberts, A. L. (2000). Pathways and kinetics of chlorinated ethylene and chlorinated acetylene
reaction with Fe(0) particles. Environmental Science and Technology, v. 34, p. 1794-1805.
Benner, S. G., Blowes, D. W., and Ptacek, C. J. (1997). A full-scale porous reactive wall for the prevention of acid mine
drainage. Ground Water Monitoring and Remediation, fall, p. 99-107.
Benner, S. G., Blowes, D. W., and Molson, J. W. H. (2001). Modeling prefential flow in reactive barriers: Implications for
performance and design. Ground Water, v. 39, p. 371-379.
Benning, L. G., Wilkin, R. T., and Barnes, H. L. (2000). Reaction pathways in the Fe-S system below 100°C. Chemical
Geology, v. 167, p. 25-51.
Berner, R. A. (1975). The role of magnesium in the crystal growth of calcite and aragonite from water. Geochimica
et Cosmochimica Acta, v. 39, 489-504.
Blengino, J. M., Keddam, M., Labbe, J. P., and Robbiola, L. (1995). Physico-chemical characterization of corrosion
layers on iron in a sodium carbonate-bicarbonate containing environment. Corrosion Science, v. 37, p. 621-643.
Blowes, D. W., Ptacek, C. J., Benner, S. G., McRae, C. W. T., Bennet, T., and Puls, R. W. (2000). Treatment of inorganic
contaminants using permeable reactive barriers. Journal of Contaminant Hydrology, v. 45, p. 123-137.
Blowes, D. W., Gillham, R. W., Ptacek, C. J., Puls, R. W., Bennett, T. A., O'Hannesin, S. F., Hanton-Fong, C. J., and
Bain, J. G. (1999a). An In Situ Permeable Reactive Barrier for the Treatment of Hexavalent Chromium and
Trichloroethylene in Ground Water, Volume 1: Design and Installation. U.S. Environmental Protection Agency, EPA/
600/R-99/095a.
Blowes, D. W., Puls, R. W., Gillham, R. W., Ptacek, C. J., Bennett, T. A., Bain, J. G., Hanton-Fong, C. J., and Paul, C.
J. (1999b). An In Situ Permeable Reactive Barrier for the Treatment of Hexavalent Chromium and Trichloroethylene
in Ground Water, Volume 2: Performance Monitoring. U.S. Environmental Protection Agency, EPA/600/R-99/095b.
Blowes, D. W. and Mayer, K. U. (1999). An In Situ Permeable Reactive Barrier for the Treatment of Hexavalent
Chromium and Trichloroethylene in Ground Water, Volume 3: Multicomponent Reactive Transport Modeling. U.S.
Environmental Protection Agency, EPA/600/R-99/095c.
Blowes, D. W., Ptacek, C. J., and Jambor, J. L. (1997). In-situ remediation of Cr(Vl)-contaminated groundwater using
permeable reactive walls: laboratory studies. Environmental Science and Technology, v. 31, p. 3348-3357.
Bonin, P. M. L., Odziemkowski, M. S., Reardon, E. J., and Gillham, R. W. (2000). In situ identification of carbonate-
containing green rust on iron electrodes in solutions simulating groundwater. Journal of Solution Chemistry, v. 29,
p. 1061-1074.
Bourrie, G., Trolard, F., Genin, J., Jaffrezic, A., Maftre, V., and Abdelmoula, M. (1999). Iron control by equilibria between
hydroxy-green rusts and solutions in hydromorphic soils. Geochimica et Cosmochimica Acta, v. 63, p. 3417-3427.
Burris, D. R., Campbell, T. J., and Manoranjan, V. S. (1995). Sorption of trichloroethylene and tetrachloroethylene in a
batch reactive metallic iron-water system. Environmental Science and Technology, v. 29, p. 2850-2855.
Butler, E. C. and Hayes, K. F. (1998). Effects of solution composition and pH on the reductive dechlorination of
hexachlorethane by iron sulfide. Environmental Science and Technology, v. 32, p. 1276-1284.
115
-------
Butler, E, C, and Hayes, K. F, (1999). Kinetics of the transformation of trichloroethylene and tetrachloroethylene by iron
sulfide. Environmental Science and Technology, v. 33, p. 2021-2027.
Butler, E. C. and Hayes, K. F. (2000). Kinetics of the transformation of halogenated aliphatic compounds by iron sulfide.
Environmental Science and Technology, v. 34, p. 422-429.
Butler, E. C. and Hayes, K. F. (2001). Factors influencing rates and products in the transformation of trichloroethylene
by iron sulfide and iron metal. Environmental Science and Technology, v. 35, p. 3884-3891.
Cantrell, K. J., Kaplan, D. I., and Wietsma, T. W. (1995). Zero-valent iron for the in situ remediation of selected metals in
groundwater. Journal of Hazardous Materials, v. 42, p. 201-212.
Chapelle, F. H., Haack, S. K., Adriaens, P., Henry, M., and Bradley, P. M. (1996). Comparison of Eh and H2
measurements for delineating redox processes in a contaminated aquifer. Environmental Science and Technology,
v. 30, p. 3565-3569.
Das, D. B. (2002). Hydrodynamic modeling for groundwater flow through permeable reactive barriers. Hydrological
Processes, v. 16, p. 3393-3418.
Davies, D. H. and Burstein, G. T. (1980). The effects of bicarbonate on the corrosion and passivation of iron. Corrosion,
v. 36, p. 416-422.
Deng, B., Burris, D. R., and Campbell, T. J. (1999). Reduction of vinyl chloride in metallic iron-water systems.
Environmental Science and Technology, v. 33, p. 2651-2656.
Dowling, J. E., Widdel, F., and White, D. C. (1986). Comparison of the phospholipid ester-linked fatty acid biomarkers of
acetate-oxidizing sulphate-reducers and other sulphide-forming bacteria. Journal of General Microbiology, v. 132, p.
1815-1825.
Drissi, S. H., Refait, Ph., Abdelmoula, M., and Genin, J. M. R. (1995). The preparation and thermodynamic properties of
Fe(ll)-Fe(lll) hydroxide-carbonate (green rust I); Pourbaix diagram of iron in carbonate-containing aqueous media.
Corrosion Science, v. 37, p. 2025-2041.
Drury, W. J. (1999). Treatment of acid mine drainage with anaerobic solid-substrate reactors. Water and Environmental
Research, v. 71, p. 1244-1250.
Dvorak, D. H., Hedin, R. S., Edenborn, H. M., and Mclntire, P. E. (1992). Treatment of metal-contaminated water using
bacterial sulfate reduction: Results from pilot scale reactors. Biotechnology and Bioengineering, v. 40, p. 609-616.
Edlund, A., Nichols, P. D., Roffey, R., and White, D. C. (1986). Extractible and lipopolysaccharide fatty acid and hydroxy
acid profiles from Desulfovibrio species. Journal of Lipid Research, v. 26, p. 982-988.
Erdos, V. and Altorfer, H. (1976). Ein dem Malachit ahnliches basisches Eisenkarbonat als Korrosionsprodukt von Stahl.
Werkstoffe und Korrosion, v. 27, p. 304-312.
Eykholt, G. R., Elder, C. R., and Benson, C. H. (1999). Effects of aquifer heterogeneity and reaction mechanism
uncertainty on a reactive barrier. Journal of Hazardous Materials, v. 68, p. 1-24.
Farrell, J., Kason, M., Melitas, N., and Li, T. (2000). Investigation of the long-term performance of zero-valent iron for
reductive dechlorination of trichloroethylene. Environmental Science and Technology, v. 34, p. 514-521.
Fennelly, J. P. and Roberts, A. L. (1998). Reaction of 1,1,1-trichloroethane with zero-valent metals and bimetallic
reductants. Environmental Science and Technology, v. 32, p. 1980-1988.
Fiedor, J. N., Bostick, W. D., Jarabek, R. J., and Farrell, J. (1998). Understanding the mechanism of uranium removal
from groundwater by zero-valent iron using X-ray photoelectron spectroscopy. Environmental Science and
Technology, v. 32, p. 1466-1473.
Furukawa, Y., Kim, J., Watkins, J., and Wilkin, R. T. (2002). Formation of ferrihydrite and associated iron corrosion
products in permeable reactive barriers of zero-valent iron. Environmental Science and Technology, v. 36, p. 5469-
5475.
Gandhi, S., Oh, B., Schnoor, J. L., and Alvarez, P. J. J. (2002). Degradation of TCE, Cr(VI), sulfate, and nitrate mixtures
by granular iron in flow-through columns under different microbial conditions. Water Research, v. 36, p. 1973-1982.
Gavaskar, A., Sass, B., Gupta, N., Drescher, E., Yoon, W., Sminchak, J., Hicks, J., and Condit, W. (2002). Evaluating
the longevity and hydraulic performance of permeable reactive barriers at Department of Defense sites. Prepared
for Naval Facilities Engineering Service Center, Port Hueneme, CA.
Gavaskar, A. R., Gupta, N., Sass, B. M., Jansoy, R. J., and O'Sullivan, D. (1998). Permeable Barriers for Groundwater
Remediation Design, Construction, and Monitoring, Battelle Press, Columbus, OH.
116
-------
Geiger, C, L, Ruiz, N. E., Clausen, C. A., Reinhart, D, R., and Quinn, J. W. (2002). Ultrasound pretreatment of elemental
iron: kinetic studies of dehalogenation reaction enhancement and surface effects. Water Research, v. 36, p. 1342-
1350.
Gillham, R. W., Blowes, D. W., Ptacek, C. J., and O'Hannesin, S. F. (1994). Use of zero-valent metals in in-situ
remediation of contaminated groundwater. In In-situ Remediation: Scientific for Current and Future
Technologies, eds. G. W. Gee and N. R. Wing, Battelle Press, Columbus, Ohio, p. 913-929.
Gillham, R. W. and O'Hannesin, S, F. (1994). Enhanced degradation of of halogenated aliphatics by zero-valent iron.
Ground Water, v. 32, p. 958-967.
Gu, B., Phelps, T. J., Liang, L., Dickey, M. J., Roh, Y., Kinsall, B. L, Palumbo, A. V., and Jacobs, G. K. (1999).
Biogeochemical dynamics in zero-valent iron columns: implications for permeable reactive barriers. Environmental
Science and Technology, v. 33, p. 2170-2177.
Gu, B., Watson, D. B., Wu, L., Phillips, D. H., White, D. C., and Zhou, J. (2002). Microbiological characterization in a
zero-valent iron reactive barrier. Environmental Monitoring and Assessment, v. 77, p. 293-307.
Gui, J. and Devine, T. M. (1994). The influence of sulfate ions on the surface enhanced Raman spectra of passive films
formed on iron. Corrosion Science, v. 36, p. 441-462.
Gupta, N. and Fox, T. C. (1999). Hydrogeologic modeling for permeable reactive barriers. Journal of Hazardous
Materials, v. 68, p. 19-39.
Hassan, S. A. (2000). Reduction of halogenated hydrocarbons in aqueous media: I. Involvement of sulfur in iron
catalysis. Chemosphere, v. 40, p. 1357-1363.
Huang, C. P., Wang, H. W., and Chiu, P. C. (1998). Nitrate reduction by metallic iron. Water Research, v. 32, p. 2257-
2264.
Hurlbut, C. S. and Klein, C. (1977). Manual of Mineralogy, 19th Edition. John Wiley & Sons, New York, 532 pp.
Johnson, T. L. and Tratnyek, P. G. (1994). A column study of geochemical factors affecting reductive dechlorination of
chlorinated solvents by zero-valent iron. In In-situ Remediation: Scientific Basis for Current and Future Technologies,
eds. G.W. Gee and N. R. Wing, Volume 2, Pacific Northwest Laboratories, Pasco, WA, p. 931-947.
Johnson, T. L., Scherer, M. M., and Tratnyek, P. G. (1996). Kinetics of halogenated organic compound degradation by
iron metal. Environmental Science and Technology, v. 30, p. 2634-2640.
Kamolpornwijit, W., Liang, L., West, O., Moline, G., and Sullivan, A. (2002). Heterogeneity development and its influence
on long-term PRB performance. In RTDF Permeable Reactive Barriers Action Team Meeting (November 6-7,
Washington, DC).
Kamolpornwijit, W., Liang, L., West, O., Moline, G., and Sullivan, A. (2003). Heterogeneity development and its influence
on long-term PRB performance: A column study. Journal of Contaminant Hydrology, in press.
Kober, R., Schlicker, O., Ebert, M., and Dahmke, A. (2002). Degradation of chlorinated ethylenes by Fe°: inhibition
process and mineral precipitation. Environmental Geology, v. 41, p. 644-652.
Lackovic, J. A., Nikolaidis, N. P., Dobbs, G. M. (2000). Inorganic arsenic removal by zero-valent iron. Environmental
Engineering and Science, v. 17, p. 29-39.
Langmuir, D. (1997). Aqueous Environmental Geochemistry. Prentice Hall, Upper Saddle River, NJ, 600 pp.
Lee, W. and Batchelor, B. (2002a). Abiotic reductive dechlorination of chlorinated ethylenes by iron-bearing soil
minerals. 1. Pyrite and magnetite. Environmental Science and Technology, v. 36, p. 5147-5154.
Lee, W. and Batchelor, B. (2002b). Abiotic reductive dechlorination of chlorinated ethylenes by iron-bearing soil
minerals. 2. Green rust. Environmental Science and Technology, v. 36, p. 5348-5354.
Liang, L., Sullivan, A., West, O., Kamolpornwijit, W., Moline, G. (2003). Predicting the precipitation of mineral phases in
permeable reactive barriers. Environmental Science and Technology, in press.
Liang, L., Moline, G., Kamolpornwijit, W., Gu, B., Sullivan, A., and West, O. (2002). Performance Assessment of Zero-
valent Iron Permeable Reactive Barriers, Environmental Sciences Division, Oak Ridge National Laboratory Report
TTP No. ORO-9-SS-30.
Liang, L., Korte, N., Gu, B., Puls, R., and Reeter, C. (2000). Geochemical and microbiological reactions affecting the
long-term performance of in situ barriers. Advances in Environmental Research, v. 4, p. 273-286.
Lien, H. and Wilkin, R. T. (2002). Reductive activation of dioxygen for degradation of methyl tert-butyl ether by
bifunctional aluminum. Environmental Science and Technology, v. 36, p. 4436-4440.
Mackenzie, P. D., Horney, D. P., and Sivavec, T. M. (1999). Mineral precipitation and porosity losses in granular iron
columns. Journal of Hazardous Materials, v. 68, p. 1-17.
117
-------
Matheson, L, J. and Tratnyek, P. G, (1994). Reductive dehalogenation of chlorinated metahnes by iron metal.
Environmental Science and Technology, v. 28, p. 2045-2053.
Matheson, L. J. (1994). Abiotic and biotic reductive dehalogenation of halogenated methanes. Unpublished Ph.D.
Dissertation, Oregon Graduate Institute of Science and Technology.
Mayer, K. U., Blowes, D. W., and Frind, E, O. (2001). Reactive transport modeling of an in situ reactive barrier for the
treatment of hexavalent chromium and trichloroethylene. Water Resources Research, v. 37, p. 3091-3103.
McGill, I. R., McEnaney, B., and Smith, D. C. (1976). Crystal structure of green rust formed by corrosion of cast iron.
Nature, v. 259, p. 200-201.
McMahon, P. B., Dennehy, K. F., and Sandstrom, M. W. (1999). Hydraulic and geochemical performance of a
permeable reactive barrier containing zero-valent iron, Denver Federal Center. Ground Water, v. 37, p. 396-404.
Melitas, N., Chuffe-Moscoso, O., and Farrell, J. (2001). Kinetics of soluble chromium removal from contaminated water
by zerovalent iron media: Corrosion inhibition and passive oxide effects. Environmental Science and Technology, v.
35, p. 3948-3953.
Moline, G., Liang, L, Kamolpornwijit, W., and West, L. (2002). Data Acquisition Report Long-term Monitoring of
Permeable Reactive Barrier, Environmental Sciences Division, Oak Ridge National Laboratory Report TTP No.
ORO-9-SS-30.
Morrison, S. J., Metzler, D. R., and Carpenter, C. E. (2001). Uranium precipitation in a permeable reactive barrier by
progressive irreversible dissolution of zerovalent iron. Environmental Science and Technology, v. 35, p. 385-390.
Morrison, S. J., Metzler, D. R., and Dwyer, B. P. (2002). Removal of As, Mn, Mo, Se, U, V and Zn from groundwater by
zero-valent iron in a passive treatment cell: reaction progress modeling. Journal of Contaminant Hydrology, v. 56, p.
99-116.
Nam, S. and Tratnyek, P. G. (2000). Reduction of azo dyes with zero-valent iron. Water Research, v. 34, p. 1837-1845.
Odziemkowski, M. S., Schumacher, T. T., and Reardon, E. J. (1998). Mechanism of oxide film formation on iron in
simulated groundwater solutions. Corrosion Science, v. 40, p. 371-389.
Oh, B., Just, C. L, and Alvarez, P. J. J. (2001). Hexahydro-1,3,5-trinitro-1,3,5-triazine mineralization by zerovalent iron
and mixed anaerobic cultures. Environmental Science and Technology, v. 35, p. 4341-4346.
O'Hannesin, S. F. and Gillham, R. W. (1998). Long-term performance of an in situ "Iron Wall" for remediation of VOCs.
Ground Water, v. 36, p. 164-172.
Orth, W. S. and Gillham, R. W. (1996). Dechlorination of trichloroethene in aqueous solution using Fe°. Environmental
Science and Technology, v. 30, p. 66-71.
Pacific Western Technologies, Ltd. 2000. 1999 Performance Evaluation of the Interim Groundwater Remediation
Measure, Denver Federal Center. Prepared for Federal Highway Administration and General Services Administration,
Denver CO.
Parkes, R. J., Dowling, N. J. E., White, D. C., Herbert, R. A., and Gibson, G. R. (1993). Characterization of sulphate-
reducing bacterial populations within marine and estuarine sediments with different rates of sulphate reduction.
FEMS Microbiology Ecology, v. 102, p. 235-250.
Phillips, D. H., Gu, B., Watson, D. B., Roh, Y., Liang, L, and Lee, S. Y. (2000). Performance evaluation of a zero-valent
iron reactive barrier: mineralogic characteristics. Environmental Science and Technology, v. 34, p. 4169-4176.
Postma, D., Boesen, C., Kristiansen, H., and Larsen, F. (1991). Nitrate reduction in an unconfined sandy aquifer: Water
chemistry, reduction processes, and geochemical modeling. Water Resources Research, v. 27, p. 255-263.
Powell, R. M., Puls, R. W., Hightower, S. K., and Sabatini, D. A. (1995). Coupled iron corrosion and chromate reduction:
Mechanisms for subsurface remediation. Environmental Science and Technology, v. 29, p. 1913-1922.
Powell, R. M., Blowes, D. W., Gillham, R. W., Schultz, D., Sivavec, T., Puls, R. W., Vogan, J. L, Powell, P. D., and
Landis, R. (1998). Permeable reactive barrier technologies for contaminant remediation, EPA/600/R-98/125, 94 pp.
Pratt, A. R., Blowes, D. W., and Ptacek, C. J. (1997). Products of chromate reduction on proposed subsurface
remediation material. Environmental Science and Technology, v. 31, p. 2492-2498.
Puls, R. W., Blowes, D. W., and Gillham, R. W. (1999a). Long-term performance monitoring for a permeable reactive
barrier at the U.S. Coast Guard Support Center, Elizabeth City, North Carolina. Journal of Hazardous Materials, v.
68, p. 109-124.
Puls, R. W., Paul, C. J., and Powell, R. M. (1999b). The application of in situ permeable reactive (zero-valent iron) barrier
technology for the remediation of chromate-contaminated groundwater: A field test. Applied Geochemistry, v. 14, p.
989-1000.
118
-------
Reardon. E, J, (1995). Anaerobic corrosion of granular iron: Measurement and interpretation of hydrogen evolution
rates. Environmental Science and Technology, v. 29, p. 2936-2945.
Refait, Ph., Gehin, A., Abdelmoula, M., and Genin, J. M. R. (2003). Coprecipitation thermodynamics of iron (ll-lll)
hydroxysulphate green rust from Fe(ll) and Fe(lll) salts. Corrosion Science, v. 45, p. 659-676.
Ritter, K., Odziemkowski, M. S., Gillham, R. W. (2002). An in situ study of the role of surface films on granular iron in the
permeable iron wall technology. Journal of Contaminant Hydrology, v. 55, p. 87-111.
Roberts, A. L, Totten, L. A., Burris, A. W., and Campbell, T. J. (1996). Reductive elimination of chlorinated ethylenes by
zero-valent metals. Environmental Science and Technology, v. 30, p. 2654-2659.
Roh, Y., Lee, S. Y., and Elless, M. P. (2000). Characterization of corrosion products in the permeable reactive barriers.
Environmental Geology, v. 40, p. 184-194.
Ruiz, N., Seal, S., and Reinhart, D. (2000). Surface chemical reactivity in selected zero-valent iron samples used in
groundwater remediation. Journal of Hazardous Materials, v. B80, p. 107-117.
Sarr, D. (2001). Zero-valent-iron permeable reactive barriers-How long will they last? Remediation, v. 11, p. 1-18.
Scherer, M. M., Johnson, K. M., Westall, J. C., and Tratnyek, P. G. (2001). Mass transport effects on the kinetics of
nitrobenzene reduction by iron metal. Environmental Science and Technology, v. 35, p. 2804-2811.
Scherer, M. M., Balko, B. A., and Tratnyek, P. G. (1998). The role of oxides in reduction reactions at the metal-water
interface. In Kinetics and Mechanisms of Reactions at the Mineral-Water Interface, Ch. 15, eds. D. L. Sparks andT.
Grundl, American Chemical Society, Washinton, D.C.
Scherer, M. M., Richter, S., Valentine, R. L., and Alvarez, P. J. (2000). Chemistry and microbiology of permeable
reactive barriers for in situ groundwater cleanup. Critical Reviews in Environmental Science and Technology, v. 30,
p. 363-411.
Schlicker, O., Ebert, M., Fruth, M., Weidner, M., WQst, W., and Dahmke, A. (2000). Degradation of TCE with iron: The
role of competing chromate and nitrate reduction. Ground Water, v. 38, p. 403-409.
Siantar, D. P., Schreier, C. G., Chou, C. S., and Reinhard, M. (1996). Treatment of 1,2-dibromo-3-chloropropane and
nitrate-contaminated water with zero-valent iron or hydrogen-palladium catalysts. Water Research, v. 30, p. 2315-
2322.
Su, C. and Puls, R.W. (1999). Kinetics of trichloroethene reduction by zerovalent iron and tin: pretreatment effect,
apparent activation energy, and intermediate products. Environmental Science and Technology, v. 33, p. 163-168.
Su, C. and Puls, R.W. (2001). Arsenate and arsenite removal by zerovalent iron: kinetics, redox transformation, and
implications for in situ groundwater remediation. Environmental Science and Technology, v. 35, p. 1487-1492.
Taylor, R. M. (1973). Crystal structures of some double hydroxide minerals. Mineralogical Magazine, v. 39, p. 377-389.
Taylor, S. W., Milly, P. C. D., and Jaffe, P. R. (1990a). Biofilm growth and the related changes in the physical properties
of a porous medium 2. Permeability. Water Resources Research, v. 26, p. 2161-2169.
Taylor, S. W. and Jaffe, P. R. (1990b). Biofilm growth and the related changes in the physical properties of a porous
medium 1. Experimental investigation. Water Resources Research, v. 26, p. 2153-2159.
Thullner, M., Mauclaire, L, Schroth, M. H., Kinzelbach, W., and Zeyer, J. (2002). Interaction between water flow and
spatial distribution of microbial growth in a two-dimensional flow field in saturated porous media. Journal of
Contaminant Hydrology, v. 58, p. 169-189.
Till, B. A., Weathers, L. J., and Alvarez, P. J. (1998). Fe(0)-supportedautotrophicdenitrification. Environmental Science
and Technology, v. 32, p. 634-639.
Tratnyek, P. G., Johnson, T. L, Scherer, M. M., and Eykholt, G. R. (1997). Remediating ground water with zero-valent
metals: Chemical considerations in barrier design. Ground Water Monitoring and Remediation, v. 17, p. 108-114.
Trudinger, P. A., Chambers, L. A., and Smith, J. W. (1985). Low-temperature sulphate reduction: Biological versus
abiological. Canadian Journal of Earth Sciences, v. 22, p. 1910-1918.
Tunlid, A. and White, D. C. (1991). Biochemical anlaysis of biomass, community structure, nutritional status and
metabolic activity of morobial communities in soil. In Soil Biochemistry, Volume 7, eds. J. M. Bollag and G. Stotzky,
New York, p.
USGS (1999). Performance evaluation of the interim ground-water remediation measure and distribution trench, Denver
Federal Center. Prepared for Federal Highway Administration and General Services Administration, Denver, CO.
Vandevivere, P. and Baveye, P. (1992). Relationship between transport of bacteria and their clogging efficiency in sand
columns. Applied and Environmental Microbiology, v. 58, p. 2523-2530.
119
-------
Vaughan, D, J. and Craig, J. R. (1978), Mineral Chemistry of Metal Sulfides. Cambridge University Press, London, 493
pp.
Vikesland, P. J., Klausen, J., Kohn, T., Burris, D. R., Ball, W. P., and Roberts, L. (2002). Co-solute effects on the
reactivity of iron with groundwater contaminants. In The Third International Conference on Remediation of
Chlorinated and Recalcitrant Compounds, Eds. A. Gavaskar and S. Chen (May 20-23, Monterey, CA).
Vogan, J. L, Focht, R. M., Clark, D. K., and Graham, S. L. (1999). Performance evaluation of a permeable reactive
barrier for remediation of dissolved chlorinated solvents in groundwater. Journal of Hazardous Materials, v. 68, p.
97-108.
Waybrant, K. R., Blowes, D. J., and Ptacek, C. J. (1998). Selection of reactive mixtures for use in permeable reactive
walls for treatment of mine drainage. Environmental Science and Technology, v. 32, p. 1972-1979.
Waybrant, K. R., Ptacek, C. J., and Blowes, D. W. (2002). Treatment of mine drainage using permeable reactive
barriers: Column experiments. Environmental Science and Technology, v. 36, p. 1349-1356.
Weathers, L. J., Parkin, G. F., and Alvarez, P. J. (1997). Utilization of cathodic hydrogen as electron donor for chloroform
cometabolism by a mixed, methanogenic culture. Environmental Science and Technology, v. 31, p. 880-885.
Wilkin, R.T., Puls, R.W., and Sewell, G.W. (2002). Long-Term Performance of Permeable Reactive Barriers for Ground
Water Remediation: An Evaluation at Two Sites. EPA Environmental Research Brief, EPA/600/S-02/001, 18 pp.
Wilkin, R.T. (2002). Permeable reactive barrier performance monitoring: Long-term trends in geochemical parameters at
two sites. In The Third International Conference on Remediation of Chlorinated and Recalcitrant Compounds, Eds.
A. Gavaskar and S. Chen (May 20-23, Monterey, CA).
Wilkin, R.T., Puls, R.W., and Sewell, G.W. (2003). Long-term performance of permeable reactive barriers using zero-
valent iron: Geochemical and microbiological effects. Ground Water, v. 41, p. 493-503.
Wolery, T.J. (1979). Calculation of chemical equilibrium between aqueous solution and minerals: The EQ3/6 software
package. Lawrence Livermore National Laboratory Report UCRL-52658.
WQst, W.F., Kober, R., Schlicker, O., and Dahmke, A. (1999). Combined zero- and first-order kinetic model of the
degradation of TCE and cis-DCE with commercial iron. Environmental Science and Technology, v. 33, p. 4304-
4309.
Yabusaki, S., Cantrell, K., Sass, B., and Steefel, C. (2001). Multicomponent reactive transport in an in situ zero-valent
iron cell. Environmental Science and Technology, v. 35, p. 1493-1503.
Zhabina, N.N. and Volkov, I.I. (1978). A method of determination of various sulfur compounds in sea sediments and
rocks. In Environmental Biogeochemistry and Geomicrobiology, W. Krumbein (ed.), v. 3, Ann Arbor, Ann Arbor
Science Publishers, p. 735-746.
120
-------
121
-------
Table A1.
Inorganic Carbon and Sulfur Concentrations in Elizabeth City Cores
Sample ID Location
Section
Interval
Date Total 1C Total S
cm M9/9 M9/9
EC050801-
EC050801-
EC050801-
EC050801-
EC050801-
EC050801-
EC050801-
EC050801-
EC050801-
EC050801-
EC050801-
EC050801-
EC050801-
EC050801-
EC050801-
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
EC050801-3 Upgradient
EC050801-3 Upgradient
EC050801-3 Upgradient
EC050801-3 Upgradient
EC050801-3 Upgradient
EC050801-3 Upgradient
EC050801-3 Upgradient
EC050801-3 Upgradient
EC050801-3 Upgradient
EC050801-3 Upgradient
EC050801-3 Upgradient
EC050801-3 Upgradient
EC050801-3 Upgradient
EC050801-3 Upgradient
EC050801-4 Upgradient
EC050801-4 Upgradient
EC050801-4 Upgradient
EC050801-4 Upgradient
EC050801-4 Upgradient
EC050801-4 Upgradient
EC050801-4 Upgradient
EC050801-4 Upgradient
EC050801-4 Upgradient
EC050801-4 Upgradient
EC050801-4 Upgradient
EC050801-4 Upgradient
EC050801-4 Upgradient
EC050801-4 Upgradient
EC050801-4 Upgradient
5
4
3
2
1
-1
-2
-3
-4
-5
-6
-7
-8
-9
-11
5
4
3
2
1
-1
-2
-3
-4
-5
-6
-7
-8
-9
5
4
3
2
1
-1
-2
-3
-4
-5
-6
-7
-8
-9
-10
0-10.2
10
20
30
40
51
56
61
66
71
76
81
86
91
.2-20.4
.4-30.6
.6-40.8
.8-51.0
.0-56.1
.1-61.2
.2-66.3
.3-71.4
.4-76.5
.5-81.6
.6-86.7
.7-91.8
.8-96.9
102-107.1
0-10.2
10
20
30
40
51
56
61
66
71
76
81
86
91
.2-20.4
.4-30.6
.6-40.8
.8-51.0
.0-56.1
.1-61.2
.2-66.3
.3-71.4
.4-76.5
.5-81.6
.6-86.7
.7-91.8
.8-99.4
0-10.2
10
20
30
40
51
56
61
66
71
76
81
86
91
96.
.2-20.4
.4-30.6
.6-40.8
.8-51.0
.0-56.1
.1-61.2
.2-66.3
.3-71.4
.4-76.5
.5-81.6
.6-86.7
.7-91.8
.8-96.9
9-107.1
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
56
31
83
121
315
1850
384
264
153
151
23
44
71
106
7
4
53
176
14
10
4633
5867
4000
2100
1086
381
453
176
149
5
7
29
20
500
1672
3337
1167
401
174
136
129
155
60
46
80
110
170
330
620
468
240
90
70
70
100
70
50
60
100
130
410
1510
190
160
3880
1100
590
340
210
150
110
80
90
60
100
140
170
690
2140
2080
570
240
216
232
285
189
1
1
122
-------
Table A1. (continued) Inorganic Carbon and Sulfur Concentrations in Elizabeth City Cores
Sample ID
Location Section Interval
Date Total 1C Total S
CIT1 UCJ/CJ LlCJ/CJ
EC050801-5
EC050801-5
EC050801-5
EC050801-5
EC050801-5
EC050801-5
EC050801-5
EC050801-5
EC050801-5
EC050801-5
EC050801-5
EC050801-6
EC050801-6
EC050801-6
EC050801-6
EC050801-6
EC050801-6
EC050801-6
EC050801-6
EC050801-6
EC050801-6
EC050801-6
EC050801-6
EC050801-7
EC050801-7
EC050801-7
EC050801-7
EC050801-7
EC050801-7
EC050801-7
EC050801-7
EC050801-7
EC050801-7
EC050801-7
EC050801-7
EC050801-7
EC050801-7
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
5
4
3
2
1
-1
-2
-3
-4
-5
-6
5
4
3
2
1
-1
-2
-3
-4
-5
-6
-7
5
4
3
2
1
-1
-2
-3
-4
-5
-6
-7
-8
-9
0-12.7
12.7-25.4
25.4-38.1
38.1-50.8
50.8-63.5
63.5-68.6
68.6-73.7
73.7-78.8
78.8-83.9
83.9-89.0
89.0-94.1
0-12.7
12.7-25.4
25.4-38.1
38.1-50.8
50.8-63.5
63.5-68.6
68.6-73.7
73.7-78.8
78.8-83.9
83.9-89.0
89.0-94.1
94.1-101.7
0-12.7
12.7-25.4
25.4-38.1
38.1-50.8
50.8-63.5
63.5-68.6
68.6-73.7
73.7-78.8
78.8-83.9
83.9-89.0
89.0-94.1
94.1-99.2
99.2-104.3
104.3-109.4
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
274
150
1023
126
258
468
4491
2370
4000
4520
1824
0
2
20
28
69
231
223
172
186
110
76
121
65
8
45
78
55
1542
3615
3636
1219
383
162
206
158
256
768
620
720
420
640
1370
1830
2680
2210
709
303
10
10
20
10
20
180
160
100
110
90
80
70
151
149
93
280
129
1610
926
278
212
168
122
95
76
147
123
-------
Table A1. (continued) Inorganic Carbon and Sulfur Concentrations in Elizabeth City Cores
Sample ID
EC050901-8
EC050901-8
EC050901-8
EC050901-8
EC050901-8
EC050901-8
EC050901-8
EC050901-8
EC050901-8
EC050901-8
EC050901-8
EC050901-9
EC050901-9
EC050901-9
EC050901-9
EC050901-9
EC050901-9
EC050901-9
EC050901-9
EC050901-9
EC050901-9
EC050901-9
EC050901-9
EC050901-9
EC060200-
EC060200-
EC060200-
EC060200-
EC060200-
EC060200-
EC060200-
EC060200-
EC060300-4
EC060300-4
EC060300-4
EC060300-4
EC060300-4
EC060300-4
EC060300-4
EC060300-4
EC060300-4
EC060300-4
EC060300-4
EC060300-4
Location
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Section
-1
-3
-5
-8
-11
-13
-15
-17
-18
-19
-20
-1
-3
-5
-7
-9
-11
-13
-15
-17
-18
-19
-20
-22
4
3
2
1
-1
-2
-3
-4
3
2
1
-1
-2
-3
-4
-5
-6
-7
-8
-9
Interval
cm
0-5.1
10.2-15.3
20.4-25.6
35.8-40.9
51.1-56.2
61 .3-66.4
71.5-76.6
81 .7-86.8
86.8-91.9
91.9-97.0
97-102.1
0-5.1
10.2-15.3
20.4-25.6
30.7-35.8
40.9-46.0
51.1-56.2
61 .3-66.4
71.5-76.6
81 .7-86.8
86.8-91.9
91.9-97.0
97.0-102.1
107.2-112.3
0-10.2
10.2-20.4
20.4-30.6
30.6-40.8
40.8-45.9
45.9-51.0
51.0-56.1
56.1-61.2
0-10.2
10.2-20.4
20.4-30.6
30.6-35.7
35.7-40.8
40.8-45.9
45.9-51.0
51.0-56.1
56.1-61.2
61.2-66.3
66.3-71.4
71 .4-76.5
Date
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
rotal 1C 1
i^g/g
59
80
51
112
49
46
90
98
201
333
285
100
122
118
204
160
242
272
276
513
517
523
854
355
7
4
3
8
308
1740
244
175
9
8
65
2339
3210
2141
234
194
119
73
72
107
fotal S
i^g/g
98
87
70
88
80
58
84
84
90
93
101
100
103
121
93
95
104
132
116
165
174
169
121
70
50
69
70
141
756
588
230
168
147
143
187
1179
466
279
301
166
161
130
104
128
124
-------
Table A1. (continued) Inorganic Carbon and Sulfur Concentrations in Elizabeth City Cores
Sample ID
Location Section
Interval
Date Total 1C Total S
CIT1 UCJ/CJ LlCJ/CJ
EC060300-6
EC060300-6
EC060300-6
EC060300-6
EC060300-6
EC060300-6
EC060300-6
EC060300-6
EC060300-6
EC060300-6
EC060300-6
EC060300-5
EC060300-5
EC060300-5
EC060300-5
EC060300-5
EC060300-5
EC060300-5
EC060300-5
EC060300-5
EC060300-5
EC060300-5
EC060500-7A
EC060500-7A
EC060500-7A
EC060500-7A
EC060500-7A
EC060500-7A
EC060500-7A
EC060500-7A
EC060500-7A
EC060500-7A
EC060500-7A
EC030616
EC030616
EC030616
EC030616
EC030616
EC030616
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
9
8
6
4
2
1
-1
-2
-3
-4
-5
8
7
6
5
4
3
2
1
-1
-2
-3
5
4
3
2
1
-1
-2
-3
-4
-5
-6
1
-1
-2
-3
-4
-5
0-7.6
7.6-15.2
22.8-30.4
38.0-45.6
53.2-60.8
60.8-68.4
68.4-73.5
73.5-78.6
78.6-83.7
83.7-88.8
88.8-93.9
0-7.6
7.6-15.2
15.2-22.8
22.8-30.4
30.4-38.0
38.0-45.6
45.6-53.2
53.2-60.8
60.8-68.4
68.4-76.0
76.0-83.6
0-7.6
7.6-15.2
15.2-22.8
22.8-30.4
30.4-38.0
38.0-43.1
43.1-48.2
48.2-53.3
53.3-58.4
58.4-63.5
63.5-68.6
0-10.2
10.2-20.4
20.4-30.6
30.6-40.8
40.8-51 .0
51.0-61.2
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
<1
<1
<1
1
41
81
92
66
81
80
242
2
10
16
12
4
10
19
7
66
57
80
na
na
na
na
na
na
na
na
na
na
na
571
1392
1345
1500
97
274
265
296
194
157
125
239
196
106
163
161
218
151
97
155
138
126
114
79
133
239
173
174
88
81
166
104
160
106
148
76
269
165
161
289
303
292
231
101
94
125
-------
Table A1. (continued) Inorganic Carbon and Sulfur Concentrations in Elizabeth City Cores
Sample ID
Location Section
Interval
Date Total 1C Total S
cm M9/9 M9/9
EC90903
EC90903
EC90903
EC90903
EC90903
EC90903
EC90903
EC90903
EC6101
EC6101
EC6101
EC6101
EC6101
EC6101
EC6101
EC6101
EC6101
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
2
1
-1
-2
-3
-4
-5
-6
-1
-2
-3
-4
-5
-6
-7
-8
-9
0-7.6
7.6-15.2
15.2-22.8
22.8-30.4
30.4-38.0
38.0-45.6
45.6-53.2
53.2-60.8
0-7.6
7.6-15.2
15.2-22.8
22.8-30.4
30.4-38.0
38.0-45.6
45.6-53.2
53.2-60.8
60.8-68.4
Sep-98
Sep-98
Sep-98
Sep-98
Sep-98
Sep-98
Sep-98
Sep-98
Jun-98
Jun-98
Jun-98
Jun-98
Jun-98
Jun-98
Jun-98
Jun-98
Jun-98
4
55
2350
357
174
117
110
50
1800
792
550
451
65
106
208
419
541
18
211
1019
315
137
183
178
142
228
195
157
117
64
96
39
34
36
126
-------
Table A2. Reduced Sulfur Speciation in Elizabeth City and Denver Federal Center Cores
Sample ID
Site
Section
Interval
Date
AVS
CRS Total S
cm M9/9 M9/9 M9/9
EC060300-4
EC060300-4
EC060300-4
EC060300-4
EC060300-4
EC060300-4
EC060300-4
EC060300-4
C2-1 2-71 300
C2-1 2-71 300
C2-1 2-71 300
C2-1 2-71 300
C2-1 2-71 300
C2-1 3-71 300
C2-1 3-71 300
C2-1 3-71 300
C2-1 3-71 300
C2-1 3-71 300
C2-1 4-71 300
C2-1 4-71 300
C2-1 4-71 300
C2-1 4-71 300
C2-1 7-71 300
C2-1 7-71 300
C2-1 7-71 300
C2-1 7-71 300
EC
EC
EC
EC
EC
EC
EC
EC
DFC
DFC
DFC
DFC
DFC
DFC
DFC
DFC
DFC
DFC
DFC
DFC
DFC
DFC
DFC
DFC
DFC
DFC
1
-1
-2
-3
-4
-6
-7
-9
-1
-3
-5
-7
-9
-1
-3
-5
-7
-9
-1
-3
-5
-7
-1
-3
-5
-7
20.4-30.6
30.6-35.7
35.7-40.8
40.8-45.9
45.9-51.0
56.1-61.2
61.2-66.3
71 .4-76.5
0-10.2
20.4-30.6
40.8-51 .0
61.2-71.4
81.6-91.8
0-10.2
20.4-30.6
40.8-51 .0
61.2-71.4
81.6-91.8
0-10.2
20.4-30.6
40.8-51 .0
61.2-71.4
0-5.1
10.2-15.3
20.4-25.5
30.6-35.7
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
17
1200
612
303
351
222
155
145
3089
4333
3019
3976
3746
3382
3434
2807
4033
3254
1538
1275
2128
1829
3590
3365
2445
1752
6
nd
39
nd
nd
nd
nd
nd
120
198
55
230
250
504
130
227
422
105
205
193
25
25
197
262
337
148
187
1179
466
279
301
161
130
128
3239
4597
3356
4084
3833
3844
3527
3172
3935
3189
1627
1519
1957
1878
3799
3366
2324
1867
Notes: nd, not determined.
127
-------
A3.
Inorganic Carbon and Sulfur Concentrations in Denver Federal Center Cores
Sample ID
Location
Section
Interval
Date
Total 1C Total S
cm M9/9 M9/9
C1-
C1-
C1-
C1-
C1-
C1-
C1-
C1-
C1-
C1-
C1-
C1-
C1-
C1-
C1-
C1-
C1-
C1-
C1-
-71701
-71701
-71701
-71701
-71701
-71701
-71701
-71701
-71701
-71701
-71701
-71701
-71701
-71701
-71701
-71701
-71701
-71701
-71701
C1 -2-71 701
C1 -2-71 701
C1 -2-71 701
C1 -2-71 701
C1 -2-71 701
C1 -2-71 701
C1 -2-71 701
C1 -2-71 701
C1 -2-71 701
C1 -2-71 701
C1 -2-71 701
C1 -3-71 701
C1 -3-71 701
C1 -3-71 701
C1 -3-71 701
C1 -3-71 701
C1 -3-71 701
C1 -3-71 701
C1 -3-71 701
C1 -3-71 701
C1 -3-71 701
C1 -3-71 701
C1 -3-71 701
C1 -3-71 701
C1 -3-71 701
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
-1
-2
-3
-4
-5
-6
-7
-8
-9
-10
-11
-12
-13
-14
-15
-16
-17
-18
-19
4
3
2
1
-1
-2
-3
-4
-5
-6
-7
3
2
1
-1
-2
-3
-4
-5
-6
-7
-8
-9
-10
-11
0-5.1
5.1-10.2
10.2-15.3
15.3-20.4
20.4-25.5
25.5-30.6
30.6-35.7
35.7-40.8
40.8-45.9
45.9-51.0
51.0-56.1
56.1-61.2
61.2-66.3
66.3-71.4
71 .4-76.5
76.5-81 .6
81 .6-86.7
86.7-91.8
91.8-102
0-10.2
10.2-20.4
20.4-30.6
30.6-40.8
40.8-45.9
45.9-51.0
51.0-56.1
56.1-61.2
61.2-66.3
66.3-71.4
71.4-79.0
0-7.6
7.6-15.2
15.2-22.8
22.8-27.9
27.9-33.0
33.0-38.1
38.1-43.2
43.2-48.3
48.3-53.4
53.4-58.5
58.5-63.6
63.6-68.7
68.7-73.8
73.8-81 .4
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
8700
2500
2200
1444
395
442
392
348
629
352
162
125
95
102
115
95
111
146
87
10300
358
31
42
9000
3700
1800
4400
2500
3407
1347
10
19
38
8
726
2600
8900
4200
3000
3000
3600
3200
2000
775
2203
2066
1966
1545
928
820
1187
874
889
899
448
463
393
359
347
326
282
242
318
390
279
72.6
151
3570
2790
1450
3160
3180
2440
1940
<5
<5
137
138
149
4321
5680
4860
4400
3290
5110
4640
2570
1570
128
-------
Table A3. (continued) Inorganic Carbon and Sulfur Concentrations in Denver Federal Center Cores
Sample ID
Location
Section
Interval
Date
Total 1C Total S
CIT1 UCJ/CJ LlCJ/CJ
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C1 -4-71 801
C2-
C2-
C2-
C2-
C2-
C2-
C2-
C2-
C2-
C2-
C2-
C2-
C2-
C2-
C2-
C2-
C2-
C2-
C2-
C2-
C2-
C2-
C2-
-71801
-71801
-71801
-71801
-71801
-71801
-71801
-71801
-71801
-71801
-71801
-71801
-71801
-71801
-71801
-71801
-71801
-71801
-71801
-71801
-71801
-71801
-71801
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
3
2
1
-1
-2
-3
-4
-5
-6
-7
-8
-9
-10
-11
-12
-13
-14
-15
-16
-17
-18
-19
2
1
-1
-2
-3
-4
-5
-6
-7
-8
-9
-10
-11
-12
-13
-14
-15
-16
-17
-18
-19
-20
-21
0-5.1
5.1-10.2
10.2-15.3
15.3-20.4
20.4-25.5
25.5-30.6
30.6-35.7
35.7-40.8
40.8-45.9
45.9-51.0
51.0-56.1
56.1-61.2
61.2-66.3
66.3-71.4
71 .4-76.5
76.5-81 .6
81 .6-86.7
86.7-91.8
91.8-96.9
96.9-102
102-107.1
107.1-112.2
0-5.1
5.1-10.2
10.2-15.3
15.3-20.4
20.4-25.5
25.5-30.6
30.6-35.7
35.7-40.8
40.8-45.9
45.9-51.0
51.0-56.1
56.1-61.2
61.2-66.3
66.3-71.4
71 .4-76.5
76.5-81 .6
81 .6-86.7
86.7-91.8
91.8-96.9
96.9-102
102-107.1
107.1-112.2
112.2-117.3
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
187
692
1500
1700
5800
9800
14100
8900
9500
11600
10400
7400
7000
3100
707
206
124
80
226
117
65
154
224
13700
7500
4700
5200
5200
2200
3000
3500
3300
3300
1800
1404
1383
2600
3300
2700
2200
1035
774
499
189
1146
144
142
237
787
2780
5290
7520
5450
5210
4430
4170
4690
3720
3010
1096
472
580
274
280
197
220
217
1580
4930
5650
4460
3950
3320
3420
2820
3550
4260
2880
2590
2090
2460
3330
3530
2270
2200
1750
1450
1210
580
2130
129
-------
Table A3. (continued) Inorganic Carbon and Sulfur Concentrations in Denver Federal Center Cores
Sample ID
Location
Section
Interval
Date
Total 1C Total S
CIT1 UCJ/CJ LlCJ/CJ
C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-4-71801
C2-4-71801
C2-4-71801
C2-4-71801
C2-4-71801
C2-4-71801
C2-4-71801
C2-4-71801
C2-4-71801
C2-4-71801
C2-4-71801
C2-4-71801
C2-4-71801
C2-4-71801
C2-4-71801
C2-4-71801
C2-4-71801
C2-4-71801
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
-1
-2
-3
-4
-5
-6
-7
-8
-9
-10
-11
-12
-13
-14
-15
-16
-17
-18
-19
2
1
-1
-2
-3
-4
-5
-6
-7
-8
-9
-10
-11
-12
-13
-14
-15
-16
0-5.1
5.1-10.2
10.2-15.3
15.3-20.4
20.4-25.5
25.5-30.6
30.6-35.7
35.7-40.8
40.8-45.9
45.9-51.0
51.0-56.1
56.1-61.2
61.2-66.3
66.3-71.4
71 .4-76.5
76.5-81 .6
81 .6-86.7
86.7-91.8
91.8-96.9
0-7.6
7.6-15.2
15.2-20.3
20.3-25.4
25.4-30.5
30.5-35.6
35.6-40.7
40.7-45.8
45.8-50.9
50.9-56.0
56.0-61.1
61.1-66.2
66.2-71.3
71.3-76.4
76.4-81 .5
81 .5-86.6
86.6-91.7
91.7-96.8
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
7700
5900
4100
2100
1324
331
137
150
128
118
91
115
90
108
122
144
111
107
135
147
4100
5900
2500
1036
398
435
296
1600
1205
1385
1498
1161
1308
1012
755
438
382
2718
2041
2111
1785
1464
582
495
561
520
506
547
300
320
440
425
501
736
637
375
<5
<5
1110
1460
758
727
611
580
1550
1250
1414
1485
1234
1189
1062
925
716
619
130
-------
Table A3. (continued) Inorganic Carbon and Sulfur Concentrations in Denver Federal Center Cores
Sample ID
Location
Section
Interval
Date
Total 1C Total S
cm M9/9 M9/9
C2- -71901
C2- -71901
C2- -71901
C2- -71901
C2- -71901
C2- -71901
C2- -71901
C2- -71901
C2- -71901
C2- -71901
C2- -71901
C2- -71901
C2- -71901
C2- -71901
C2- -71901
C2-2-71901
C2-2-71901
C2-2-71901
C2-2-71901
C2-2-71901
C2-2-71901
C2-2-71901
C2-2-71901
C2-2-71901
C2-2-71901
C2-2-71901
C2-2-71901
C2-2-71901
C2-2-71901
C2-3-71901
C2-3-71901
C2-3-71901
C2-3-71901
C2-3-71901
C2-3-71901
C2-3-71901
C2-3-71901
C2-3-71901
C2-3-71901
C2-3-71901
C2-3-71901
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
-1
-2
-3
-4
-5
-6
-7
-8
-9
-10
-11
-12
-13
-14
-15
-1
-2
-3
-4
-5
-6
-7
-8
-9
-10
-11
-12
-13
-14
1
2
-1
-2
-3
-4
-5
-6
-7
-8
-9
-10
0-7.6
7.6-15.
15.2-22
22.8-30
30.4-38
38.0-45
45.6-53
53.2-60
60.8-68
68.4-76
76.0-83
83.6-91
91.2-98
2
.8
.4
.0
.6
.2
.8
.4
.0
.6
.2
.8
98.8-106.4
106.4-114.0
0-7.6
7.6-15.
15.2-22
22.8-30
30.4-38
38.0-45
45.6-53
53.2-60
60.8-68
68.4-76
76.0-83
83.6-91
91.2-98
2
.8
.4
.0
.6
.2
.8
.4
.0
.6
.2
.8
98.8-106.4
0-7.6
7.6-15.
15.2-22
22.8-30
30.4-38
38.0-45
45.6-53
53.2-60
60.8-68
68.4-76
76.0-83
83.6-91
2
.8
.4
.0
.6
.2
.8
.4
.0
.6
.2
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
417
567
384
960
1203
1523
1116
1349
1283
2600
2050
2850
2700
2600
4500
212
184
217
249
365
352
1700
253
291
176
172
124
216
146
2500
7900
1299
1113
2500
5600
4500
4100
2800
3000
4000
3300
920
1192
1220
1742
2150
2807
2621
1879
2291
2995
3108
3175
3803
2493
3374
814
755
546
628
499
380
613
633
210
694
551
687
232
782
122
269
831
971
1590
1910
1750
1660
1730
1840
1490
1420
131
-------
Table A3. (continued) Inorganic Carbon and Sulfur Concentrations in Denver Federal Center Cores
Sample ID
Location Section
Interval
Date Total 1C Total S
cm M9/9 M9/9
C3-2-71801
C3-2-71801
C3-2-71801
C3-2-71801
C3-2-71801
C3-2-71801
C3-2-71801
C3-2-71801
C3-2-71801
C3-2-71801
C3-2-71801
C3-2-71801
C3-2-71801
C3-2-71801
C3-2-71801
C3-2-71801
C1 -2-71 000
C1 -2-71 000
C1 -2-71 000
C1 -2-71 000
C1 -2-71 000
C1 -2-71 000
C1 -2-71 000
C1 -2-71 000
C1 -2-71 000
C1 -2-71 000
C1 -2-71 000
C1 -2-71 000
C1 -2-71 000
C1 -2-71 000
C1 -2-71 000
C1 -3-71 100
C1 -3-71 100
C1 -3-71 100
C1 -3-71 100
C1 -3-71 100
C1 -3-71 100
C1 -3-71 100
C1 -3-71 100
C1 -3-71 100
C1 -3-71 100
C1 -3-71 100
C1 -3-71 100
C1 -3-71 100
C1 -3-71 100
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
-1
-2
-3
-4
-5
-6
-7
-8
-9
-10
-11
-12
-13
-14
-15
-16
-1
-2
-3
-4
-5
-6
-7
-8
-9
-10
-11
-12
-13
-14
-15
3
2
1
-1
-2
-3
-4
-5
-6
-7
-8
-9
-10
-11
0-5.1
5.1-10.2
10.2-15.3
15.3-20.4
20.4-25.5
25.5-30.6
30.6-35.7
35.7-40.8
40.8-45.9
45.9-51.0
51.0-56.1
56.1-61.2
61.2-66.3
66.3-71.4
71 .4-76.5
76.5-81 .6
0-5.1
5.1-10.2
10.2-15.3
15.3-20.4
20.4-25.5
25.5-30.6
30.6-35.7
35.7-40.8
40.8-45.9
45.9-51.0
51.0-56.1
56.1-61.2
61.2-66.3
66.3-73.9
73.9-81.5
0-7.6
7.6-15.2
15.2-22.8
22.8-27.9
27.9-33.0
33.0-38.1
38.1-43.2
43.2-48.3
48.3-53.4
53.4-58.5
58.5-63.6
63.6-68.7
68.7-73.8
73.8-81 .4
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-01
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
1800
594
452
429
693
1037
1900
640
1200
2000
2800
3900
6300
4900
4100
5800
593
854
1390
1559
1056
563
493
361
202
306
235
118
144
125
167
2656
906
769
245
218
161
154
158
160
174
126
139
128
143
231
203
152
133
146
149
881
401
547
649
1180
1940
2320
2290
1051
2120
702
1141
1377
1328
1107
681
648
437
431
460
421
306
267
278
284
4143
2518
1199
717
464
412
413
303
232
233
169
128
143
106
132
-------
Table A3. (continued) Inorganic Carbon and Sulfur Concentrations in Denver Federal Center Cores
Sample ID
Location Section
Interval
Date Total 1C Total S
CIT1 UCJ/CJ LlCJ/CJ
C2-1 7-71 300
C2-1 7-71 300
C2-1 7-71 300
C2-1 7-71 300
C2-1 7-71 300
C2-1 7-71 300
C2-1 7-71 300
C2-1 7-71 300
C2-1 6-71 300
C2-1 6-71 300
C2-1 6-71 300
C2-1 6-71 300
C2-1 6-71 300
C2-1 2-71 300
C2-1 2-71 300
C2-1 2-71 300
C2-1 2-71 300
C2-1 2-71 300
C2-1 2-71 300
C2-1 2-71 300
C2-1 2-71 300
C2-1 2-71 300
C2-1 2-71 300
C2-1 3-71 300
C2-1 3-71 300
C2-1 3-71 300
C2-1 3-71 300
C2-1 3-71 300
C2-1 3-71 300
C2-1 3-71 300
C2-1 3-71 300
C2-1 3-71 300
C2-1 4-71 300
C2-1 4-71 300
C2-1 4-71 300
C2-1 4-71 300
C2-1 4-71 300
C2-1 4-71 300
C2-1 4-71 300
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Upgradient
Downgradient
Downgradient
Downgradient
Downgradient
Downgradient
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
Vertical
-1
-2
-3
-4
-5
-6
-7
-8
1
-1
-2
-3
-4
-1
-2
-3
-4
-5
-6
-7
-8
-9
-10
-1
-2
-3
-4
-5
-6
-7
-8
-9
-1
-2
-3
-4
-5
-6
-7
0-5.1
5.1-10.2
10.2-15.3
15.3-20.4
20.4-25.5
25.5-30.6
30.6-35.7
35.7-43.3
0-7.6
7.6-12.7
12.7-17.8
17.8-22.9
22.9-28.0
0-10.2
10.2-20.4
20.4-30.6
30.6-40.8
40.8-51 .0
51.0-61.2
61.2-71.4
71.4-81.6
81.6-91.8
91.8-102
0-10.2
10.2-20.4
20.4-30.6
30.6-40.8
40.8-51 .0
51.0-61.2
61.2-71.4
71.4-81.6
81.6-91.8
0-10.2
10.2-20.4
20.4-30.6
30.6-40.8
40.8-51 .0
51.0-61.2
61.2-71.4
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
5987
4324
4772
4817
3704
3481
2367
2769
871
793
2299
4243
5949
5470
6048
7233
7201
7391
7506
4613
6556
6692
8969
5178
6786
6602
5603
4035
3850
7058
5073
5637
1523
476
799
1360
2495
3577
3284
3799
3546
3366
2838
2324
2092
1867
1678
913
1320
2034
3830
3362
3239
3966
4597
3886
3356
3493
4084
4453
3833
2044
3844
3761
3527
2923
3172
3354
3935
3324
3189
1627
1207
1519
1474
1957
1883
1878
133
-------
134
-------
B
135
-------
Table Bl. Fatty-acid (PLFA) from City
Sample ID
EC050801-5
EC050801-5
EC050801-5
EC050801-5
EC050801-5
EC050801-5
EC050801-5
EC050801-5
EC050801-5
EC050801-6
EC050801-6
EC050801-6
EC050801-6
EC050801-6
EC050801-6
EC050801-6
EC050801-6
EC050801-6
EC050801-6
EC050801-7
EC050801-7
EC050801-7
EC050801-7
EC050801-7
EC050801-7
EC050801-7
EC050801-7
EC050801-7
EC060200-1
EC060200-1
EC060200-1
EC060200-1
EC060200-1
EC060200-1
EC060200-1
EC060200-1
EC060200-4
EC060200-4
EC060200-4
EC060200-4
EC060200-4
EC060200-4
EC060200-4
EC060200-4
EC060200-4
EC060200-4
EC060200-4
EC060200-4
Section Date
3
2
1
-1
-2
-3
-4
-5
-6
3
2
1
-1
-2
-3
-4
-5
-6
-7
3
2
1
-1
-2
-3
-4
-7
-9
4
3
2
1
-1
-2
-3
-4
3
2
1
-1
-2
-3
-4
-5
-6
-7
-8
-9
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
May-01
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
PLFA
pM/g
155
248
348
782
2614
1809
2406
260
24
27
8
5
54
9
5
6
4
3
ND
25
5
18
913
196
25
9
6
6
4
3
9
89
875
309
44
51
4
5
51
570
203
173
41
13
4
3
42
5
PLFA
cells/g
3.11E+06
4.95E+06
6.96E+06
1.56E+07
5.23E+07
3.62E+07
4.81E+07
5.20E+06
4.72E+05
5.38E+05
1.66E+05
1.06E+05
1.07E+06
1.86E+05
1.09E+05
1.24E+05
8.77E+04
6.95E+04
ND
5.05E+05
9.28E+04
3.56E+05
1.83E+07
3.92E+06
5.00E+05
1.77E+05
1.19E+05
1.15E+05
7.10E+04
5.44E+04
1.87E+05
1.78E+06
1.75E+07
6.18E+06
8.73E+05
1.02E+06
7.56E+04
9.52E+04
1.02E+06
1.14E+07
4.07E+06
3.46E+06
8.24E+05
2.67E+05
7.73E+04
6.55E+04
8.46E+05
9.63E+04
Biomass
Prokaryote
PLFA
pM/g
154
245
3341
782
2614
1809
2406
260
24
27
8
5
50
9.3
5.4
6.2
4.4
3.5
ND
25
5
17
913
196
25
9
6
6
3
3
9
88
875
309
44
51
4
5
51
569
203
173
41
13
4
3
42
5
Eukaryotic
PLFA
pM/g
1
2
7
ND
ND
ND
ND
ND
ND
ND
ND
ND
4
ND
ND
ND
ND
ND
ND
ND
ND
1
ND
ND
ND
ND
ND
ND
Trace
ND
ND
1
ND
1
Trace
Trace
ND
Trace
Trace
1
Trace
Trace
ND
ND
ND
ND
Trace
ND
Prokaryote/
Eukaryote
174
100
52
NO
NO
NO
NO
NO
NO
NO
NO
NO
13
NO
NO
NO
NO
NO
NO
NO
NO
16
NO
NO
NO
NO
NO
NO
17
19
NO
148
NO
555
832
178
NO
27
138
434
525
666
NO
NO
NO
NO
249
NO
Gram+/
Anaerobic Gram-
(TerBrSats)
19.2
17.7
18.3
28.3
42.1
38.1
41.1
28.3
21.9
0.0
0.0
0.0
12.8
0.0
0.0
0.0
0.0
0.0
0.0
16.5
0.0
10.1
18.1
16.0
14.1
5.3
0.0
0.0
0.0
0.0
0.0
5.8
7.4
12.6
7.5
3.8
0.0
8.1
4.5
24.8
32.5
27.7
22.0
12.4
13.2
2.7
2.2
16.8
Community Structure (% of total PLFA)
Gram- Anaerobic SRB/ Genera
(Monos) Metal Reducers Actinomycetes Nsats
27.7
30.6
28.9
29.4
17.1
19.0
18.8
20.9
33.1
0.0
59.9
52.6
41.9
69.0
69.0
71.5
50.8
53.8
0.0
30.6
22.2
37.8
31.3
35.2
45.0
73.5
74.8
58.9
34.2
29.8
34.7
37.2
57.7
50.9
30.1
16.5
37.0
34.7
15.1
17.6
17.4
14.4
20.4
35.8
41.4
22.7
10.1
30.7
(BrMonos)
28.2
26.1
27.2
23.6
30.3
29.3
28.5
37.8
27.9
0.0
0.0
0.0
16.6
0.0
0.0
0.0
0.0
0.0
0.0
26.4
29.8
20.7
26.5
27.4
20.7
0.0
0.0
0.0
0.0
0.0
0.0
0.5
3.6
12.7
4.8
2.5
0.0
0.0
0.4
8.0
12.7
13.6
12.8
4.8
2.6
0.0
0.4
0.0
(MidBrSats)
6.9
7.0
5.5
0.4
0.0
2.8
0.0
0.9
0.0
0.0
0.0
0.0
2.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
1.1
3.3
0.0
0.0
0.0
6.3
7.0
0.0
6.0
4.0
9.5
3.9
2.2
0.0
12.3
3.7
18.5
30.3
38.9
28.9
7.9
4.1
0.0
1.0
0.0
17.5
17.7
18.3
18.4
10.6
10.9
11.7
12.1
17.2
100.0
40.1
47.4
19.4
31.0
31.1
28.5
49.3
46.2
0.0
26.6
48.1
25.6
24.2
20.4
16.9
21.2
25.2
41.1
53.9
58.2
65.3
49.9
27.3
14.2
53.6
74.5
63.0
41.3
75.6
30.9
7.0
5.2
15.9
39.1
38.8
74.6
85.9
52.6
Eukaryotes
(Polyenoics)
0.6
1.0
1.9
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
7.2
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
5.8
0.0
0.0
0.0
0.0
0.0
0.0
5.7
5.0
0.0
0.7
0.0
0.2
0.1
0.6
0.0
3.6
0.7
0.2
0.2
0.2
0.0
0.0
0.0
0.0
0.4
0.0
Physiological
Status
0.00
0.00
0.00
0.03
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.10
0.10
0.20
0.04
0.18
0.00
0.00
0.00
0.38
0.94
0.85
0.51
0.27
0.00
0.00
0.00
0.00
-------
Table Bl. Fatty-acid (PLFA) from City
Sample ID
EC060300-5
EC060300-5
EC060300-5
EC060300-5
EC060300-5
EC060300-5
EC060300-5
EC060300-5
EC060300-5
EC060300-5
EC060300-5
EC060300-6
EC060300-6
EC060300-6
EC060300-6
EC060300-6
EC060300-6
EC060300-6
EC060300-6
EC060300-6
EC060300-6
_^ EC060300-6
CO EC060300-6
~"J EC060300-6
EC060300-6
EC90902
EC90902
EC90902
EC90902
EC90902
EC90902
EC90902
EC90902
EC90903
EC90903
EC90903
EC90903
EC90903
EC90903
EC90903
EC90903
EC01061899
EC01061899
EC01061899
EC03061699
EC03061699
EC03061699
EC03061699
Section Date
8
7
6
5
4
3
2
1
-1
-2
-3
9
8
7
6
5
4
3
2
1
-1
-2
-3
-4
-5
8
7
6
5
4
3
2
1
2
1
-1
-2
-3
-4
-5
-6
-1
-2
-3
-1
-2
-4
-5
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Jun-00
Sep-99
Sep-99
Sep-99
Sep-99
Sep-99
Sep-99
Sep-99
Sep-99
Sep-99
Sep-99
Sep-99
Sep-99
Sep-99
Sep-99
Sep-99
Sep-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
PLFA
pM/g
4
12
2
12
11
9
6
57
14
7
21
22
41
1
0
0
1
6
54
5
15
25
9
6
92
5
1
1
2
1
1
3
4
51
219
2610
114
53
22
11
15
128
39
38
474
1171
130
69
PLFA
cells/g
8.51E+04
2.33E+05
4.18E+04
2.40E+05
2.12E+05
1.80E+05
1.20E+05
1.13E+06
2.79E+05
1.42E+05
4.25E+05
4.39E+05
8.13E+05
2.98E+04
2.51E+03
6.92E+03
1.88E+04
1.14E+05
1.07E+06
1.09E+05
3.06E+05
4.91E+05
1.82E+05
1.23E+05
1.85E+06
9.19E+04
2.99E+04
1.93E+04
4.19E+04
1.88E+04
1.77E+04
5.79E+04
7.77E+04
1.02E+06
4.38E+06
5.22E+07
2.28E+06
1.06E+06
4.40E+05
2.26E+05
3.03E+05
2.55E+06
7.82E+05
7.54E+05
9.48E+06
2.34E+07
2.60E+06
1.38E+06
Biomass
Prokaiyote
PLFA
pM/g
4
11
2
12
10
9
6
56
14
7
21
22
40
1
0
0
1
6
54
5
15
24
9
6
92
5
1
1
2
1
1
3
4
559
1541
2610
114
53
22
11
15
125
38
33
463
1159
61
30
Eukaiyotic
PLFA
pM/g
Trace
Trace
Trace
Trace
1
Trace
Trace
1
Trace
ND
ND
Trace
1
ND
ND
ND
ND
ND
ND
ND
Trace
Trace
Trace
ND
1
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
3
1
3
11
11
64
37
Prokaryote/
Eukaryote
14
43
10
38
20
34
24
59
146
NO
NO
56
64
NO
NO
NO
NO
NO
NO
NO
75
121
50
NO
131
NO
NO
NO
NO
NO
NO
NO
NO
NO
NO
NO
NO
NO
NO
NO
NO
47
34
12
41
101
1
1
Gram+/
Anaerobic Gram-
(TerBrSats)
4.1
8.6
0.0
4.2
2.6
9.8
16.7
11.3
3.1
0.0
0.6
2.8
2.7
0.0
0.0
0.0
0.0
0.0
12.0
5.2
0.0
1.7
0.0
0.0
2.1
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
13.9
15.6
40.1
22.6
10.1
21.1
0.0
13.6
15.8
4.5
2.2
20.2
25.5
1.7
2.1
Community Structure (% of total PLFA)
Gram- Anaerobic SRB/ Genera
(Monos) Metal Reducers Actinomycetes Nsats
21.8
18.1
21.3
26.1
12.1
11.1
20.9
16.0
13.7
23.5
12.5
25.7
17.2
0.0
0.0
0.0
0.0
17.2
63.0
25.0
14.1
32.7
33.1
17.5
13.3
50.9
50.1
33.6
46.6
30.4
39.7
38.9
47.7
38.7
40.0
23.5
38.4
63.3
50.9
66.9
60.0
30.6
10.3
5.5
36.3
32.4
14.0
10.0
(BrMonos)
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.4
0.0
8.2
0.0
0.0
2.1
1.3
0.0
0.0
0.0
2.8
5.2
0.0
0.6
0.0
0.0
0.0
1.5
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
6.3
2.2
6.8
3.3
2.8
2.4
0.0
1.2
12.3
3.1
1.3
11.7
14.2
1.0
1.2
(MidBrSats)
0.0
0.0
0.0
0.8
2.6
2.7
7.2
3.5
0.4
0.0
0.0
0.0
9.1
3.0
0.0
0.0
0.0
0.0
4.2
0.0
0.0
0.0
0.0
0.0
1.5
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
17.1
8.2
17.3
6.3
1.2
0.0
0.0
0.0
8.1
2.3
1.1
11.6
14.1
1.0
1.4
67.4
71.0
69.2
66.3
78.0
73.6
51.2
67.2
82.1
68.4
86.9
58.6
74.2
100.0
100.0
100.0
100.0
80.1
15.6
69.8
84.6
64.2
64.9
82.5
80.9
49.1
49.9
66.4
53.4
69.6
60.3
61.2
52.3
24.1
34.2
12.4
29.5
22.7
25.6
33.1
25.3
31.1
75.8
77.6
17.9
12.8
29.5
28.4
Eukaryotes
(Polyenoics)
6.8
2.3
9.5
2.6
4.7
2.9
4.1
1.7
0.7
0.0
0.0
1.8
1.6
0.0
0.0
0.0
0.0
0.0
0.0
0.0
1.3
0.8
2.0
0.0
0.8
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
2.1
2.8
7.5
2.4
1.0
49.6
53.7
Physiological
Status
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.13
0.00
0.00
0.00
0.20
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.00
0.10
0.19
0.00
0.26
0.00
0.00
NO
0.00
NO
NO
0.00
0.00
0.39
0.13
NO
0.30
0.38
0.00
0.00
0.00
0.07
0.10
0.00
0.07
0.07
0.00
0.00
-------
Table B1. Fatty-acid (PLFA) from City
03
O3
Sample ID
EC6102
EC6102
EC6102
EC6102
EC6101
EC6101
EC6101
EC6101
EC6101
EC6101
EC6101
EC6101
EC6101
EC6101
EC6114
EC6114
EC6114
EC6114
EC6114
EC6114
EC6114
EC6114
Section
4
3
2
1
-1
-2
-3
-4
-5
-6
-7
-8
-9
-10
4
3
2
1
-1
-2
-3
-4
Date
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
Jun-99
PLFA
pM/g
1
3
1
2
795
85
14
6
5
2
1
3
1
96
412
39
8
6
21
8
2
3
PLFA
cells/g
1.99E+04
5.21E+04
1.32E+04
3.06E+04
1.59E+07
1.70E+06
2.72E+05
1.26E+05
1.05E+05
3.59E+04
1.25E+05
6.90E+04
2.03E+04
1.92E+06
8.24E+06
7.84E+05
1.62E+05
1.17E+05
4.21E+05
1.59E+05
3.27E+04
6.11E+04
Biomass
Prokaryote
PLFA
pM/g
1
3
1
2
795
85
14
6
5
2
1
3
1
70
412
39
8
6
21
8
2
3
Eukaryotic
PLFA
pM/g
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
26
ND
ND
ND
ND
ND
ND
ND
ND
Prokaryote/
Eukaryote
NO
NO
NO
NO
NO
NO
NO
NO
NO
NO
NO
NO
NO
3
NO
NO
NO
NO
NO
NO
NO
NO
Gram+/
Anaerobic Gram-
(TerBrSats)
0.0
0.0
0.0
0.0
15.7
9.9
7.1
0.0
0.0
0.0
0.0
0.0
0.0
1.4
27.5
11.5
21.1
0.0
3.3
1.6
0.0
14.4
Community Structure (% of total PLFA)
Gram- Anaerobic SRB/ Genera
(Monos) Metal Reducers Actinomycetes Nsats
31.0
45.9
57.0
0.0
39.8
59.5
77.2
66.9
65.9
29.3
0.0
37.5
18.1
51.9
37.9
56.1
54.4
58.1
58.5
31.6
0.0
18.6
(BrMonos)
1.0
1.0
1.0
0.0
7.0
4.8
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
7.6
4.2
0.0
0.0
0.0
0.0
0.0
0.0
(MidBrSats)
0.0
0.0
0.0
0.0
7.9
6.5
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.1
12.1
5.4
0.0
0.0
2.2
7.4
0.0
0.0
69.0
54.1
43.0
100.0
29.7
19.2
15.7
33.1
34.1
70.7
100.0
62.5
81.9
19.4
14.9
22.8
24.4
41.9
36.0
59.5
100.0
66.9
Eukaryotes
(Polyenoics)
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
27.3
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
Physiological
Status
NO
NO
NO
NO
0.17
0.14
0.00
0.00
NO
NO
0.00
0.00
0.00
0.30
0.10
0.00
0.00
0.00
0.20
0.00
NO
NO
The number of cells is calculated based on 2.0x1012 cells per gram dry weight of cells and 10s picolmoles of phospholipid per gram dry weight of cells. Microbial Insights (G. Davis, pers. comm.)
points out that the number of cells/gram of dry weight may vary and is dependent on the environmental conditions from which the microorganisms were recovered. Physiological status is based on ratios
of fatty acid biomarkers that indicate a metabolic response to environmentally induced stress evidenced by decreased membrane permeability (w7/w7c). The biomarker indicators generally suggest that the samples
have slow to moderate rates of turnover. Sample locations are shown in Figure 3.19. ND, not detected; NC not calculated.
-------
Table B2. Fatty-acid (PLFA) from
Sample ID
C1-1-71701
C1-1-71701
C1-1-71701
C1-1-71701
C1-1-71701
C1-1-71701
C1-1-71701
C1-1-71701
C2-1-71801
C2-1-71801
C2-1-71801
C2-1-71801
C2-1-71801
C2-1-71801
C2-1-71801
C2-1-71801
C2-1-71801
C2-1-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
_^ C2-3-71801
CO C2-3-71801
10 C2-3-71801
C2-3-71801
C2-3-71801
C2-3-71801
C2-1-71901
C2-1-71901
C2-1-71901
C2-1-71901
C2-1-71901
C2-1-71901
C2-2-71901
C2-2-71901
C2-2-71901
C2-2-71901
C2-2-71901
C2-3-71901
C2-3-71901
C2-3-71901
C2-3-71901
C3-2-71801
C3-2-71801
C3-2-71801
C3-2-71801
C3-2-71801
Section
-1
-2
-3
-4
-6
-8
-12
-18
2
-1
-2
-3
-4
-5
-7
-11
-15
-19
-1
-2
-3
-4
-6
-9
-12
-15
-17
-19
-2
-4
-6
-10
-12
-14
-2
-6
-9
-12
-14
-1
-2
-6
-9
-5
-10
-12
-14
-16
Date
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
Jul-02
PLFA
pM/g
1533
979
775
858
222
127
27
22
68
1439
1091
1100
1561
1517
1556
984
1372
332
1869
773
533
463
125
111
58
61
114
50
363
1326
1075
1909
2874
1527
42
24
60
42
86
208
100
337
656
234
655
977
1439
1301
Biomass
Prokaiyote
PLFA PLFA
cells/g
3.07E+07
1.96E+07
1.55E+07
1.72E+07
4.43E+06
2.55E+06
5.42E+05
4.33E+05
1.36E+06
2.88E+07
2.18E+07
2.20E+07
3.12E+07
3.03E+07
3.11E+07
1.97E+07
2.74E+07
6.63E+06
3.74E+07
1.55E+07
1.07E+07
9.25E+06
2.51E+06
2.22E+06
1.17E+06
1.23E+06
2.29E+06
1.00E+06
7.26E+06
2.65E+07
2.15E+07
3.82E+07
5.75E+07
3.05E+07
8.50E+05
4.73E+05
1.20E+06
8.38E+05
1.72E+06
4.16E+06
2.00E+06
6.74E+06
1.31E+07
4.68E+06
1.31E+07
1.95E+07
2.88E+07
2.60E+07
pM/g
1499
954
758
838
213
123
26
20
66
1410
1066
1084
1527
1477
1533
970
1348
324
1853
767
526
455
121
105
58
59
112
49
345
1303
1053
1883
2835
1497
39
19
55
39
85
191
94
331
617
220
638
965
1431
1294
Eukaryotic
PLFA
pM/g
34
25
17
20
9
4
1
2
1
29
25
17
34
40
24
14
24
8
16
6
6
8
4
6
1
3
2
1
17
23
22
25
38
30
3
4
5
3
1
17
6
6
38
14
17
12
8
7
Prokaiyote/
Eukaryote
44
38
44
42
24
28
28
10
46
50
42
67
45
37
64
69
57
43
119
124
82
60
27
18
68
21
47
52
20
56
48
74
74
49
13
4
12
14
59
11
16
57
16
16
37
82
178
195
Gram+/
Anaerobic Gram-
(TerBrSats)
9.3
8.1
10.0
11.4
9.1
6.5
4.0
2.5
8.7
17.1
18.3
20.5
17.8
17.6
13.8
14.2
14.1
11.9
26.4
22.1
20.3
17.2
8.5
5.5
3.4
4.0
6.8
7.2
9.3
20.3
15.3
19.4
20.2
15.5
4.0
3.1
3.6
4.3
14.0
17.8
10.1
17.8
17.8
13.8
18.4
21.4
22.5
22.3
Community Structure (% of total PLFA)
Gram- Anaerobic SRB/ Genera
(Monos) Metal Reducers Actinomycetes Nsats
51.6
55.7
51.5
47.1
53.5
60.0
72.2
71.0
51.9
41.4
43.2
42.2
47.8
46.9
55.3
55.8
54.1
54.9
31.6
39.4
44.3
51.5
72.2
56.3
85.3
78.2
81.2
79.1
62.7
42.4
43.9
36.1
41.2
44.1
77.7
61.7
76.1
74.9
72.5
49.6
67.0
32.3
38.5
52.4
47.0
34.0
30.6
30.9
(BrMonos)
9.8
7.9
8.0
10.9
10.7
8.1
4.3
0.0
10.3
12.9
12.2
12.4
10.4
11.3
8.9
9.6
7.8
8.4
12.2
11.9
10.3
8.6
3.0
1.8
0.0
1.1
1.0
1.3
9.4
15.2
14.8
16.5
15.3
15.4
1.0
3.6
0.0
1.1
1.4
2.6
3.7
9.6
9.9
11.7
12.8
15.8
15.9
14.5
(MidBrSats)
6.2
4.8
4.8
6.0
5.9
4.7
0.0
0.0
8.5
10.9
9.0
8.3
6.0
6.4
5.0
5.2
4.3
10.6
9.4
7.9
6.5
5.6
2.0
21.2
0.0
0.0
0.0
0.0
4.4
6.2
9.1
10.0
4.6
8.9
0.0
0.0
3.7
0.0
0.0
2.2
2.6
8.2
14.3
3.4
5.6
8.9
7.4
5.3
20.9
21.0
23.4
22.2
16.8
17.3
16.0
17.1
18.5
15.8
15.1
15.0
15.8
15.2
15.6
13.8
18.0
11.8
19.7
17.9
17.4
15.4
10.9
10.1
9.9
12.2
8.9
10.6
9.4
14.3
14.9
16.7
17.4
14.1
10.2
13.2
8.6
12.9
10.5
19.6
10.8
30.6
13.8
12.9
13.6
18.7
23.0
26.6
Eukaryotes
(Polyenoics)
2.2
2.6
2.2
2.3
4.1
3.5
3.5
9.4
2.1
2.0
2.3
1.5
2.2
2.6
1.5
1.4
1.7
2.3
0.8
0.8
1.2
1.6
3.5
5.1
1.5
4.6
2.1
1.9
4.8
1.7
2.1
1.3
1.3
2.0
7.2
18.4
8.0
6.9
1.7
8.2
5.8
1.7
5.8
5.8
2.6
1.2
0.6
0.5
Physiological
Status
0.26
0.26
0.24
0.21
0.17
0.17
0.00
0.00
0.16
0.42
0.32
0.32
0.27
0.28
0.26
0.15
0.24
0.13
0.25
0.23
0.18
0.19
0.05
0.06
0.04
0.05
0.04
0.06
0.06
0.34
0.13
0.17
0.46
0.21
0.00
0.00
0.05
0.06
0.16
0.12
0.08
0.17
0.15
0.32
0.44
0.56
0.50
0.36
-------
Table B2. Fatty-acid (PLFA) from
Sample ID
C1-2-71000
C1-2-71000
C1-2-71000
C1-2-71000
C1-2-71000
C1-2-71000
C1-2-71000
C1-3-71100
C1-3-71100
C1-3-71100
C1-3-71100
C1-3-71100
C2-1 3-71 300
C2-1 3-71 300
C2-1 3-71 300
C2-1 3-71 300
C2-1 3-71 300
C2-1 4-71 300
C2-1 4-71 300
C2-1 4-71 300
C2-1 4-71 300
C2-1 7-71 300
C2-1 7-71 300
C2-1 7-71 300
C2-1 7-71 300
C2-1 7-71 300
C1-2-52199
C1-3-52199
C1-3-52199
C1-3-52199
C2-1 -52299
C2-1 -52299
C2-1 -52299
Section
-1
-2
-3
-4
-5
-8
-13
3
1
-1
-2
-4
-1
-3
-5
-7
-9
-1
-3
-5
-7
-1
-2
-3
-5
-7
1
-1
-3
-4
4
-1
-2
Date
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
Jul-00
May-99
May-99
May-99
May-99
May-99
May-99
May-99
PLFA
pM/g
1904
95
693
506
103
15
22
124
306
80
19
30
2752
4182
1827
1115
2296
601
576
283
916
3851
3760
4066
3916
1820
592
2584
3554
3455
1629
4523
4924
Biomass
Prokaiyote
PLFA PLFA
cells/g
3.81E+07
1.89E+06
1.39E+07
1.01E+07
2.07E+06
3.03E+05
4.48E+05
2.49E+06
6.13E+06
1.59E+06
3.85E+05
6.00E+05
5.50E+07
8.36E+07
3.65E+07
2.23E+07
4.59E+07
1.20E+07
1.15E+07
5.66E+06
1.83E+07
7.70E+07
7.52E+07
8.13E+07
7.83E+07
3.64E+07
1.18E+07
5.17E+07
7.11E+07
6.91E+07
3.26E+07
9.05E+07
9.85E+07
pM/g
1776
93
691
505
103
15
22
122
304
79
19
29
2752
4177
1821
1115
2297
599
574
282
913
3842
3752
4061
3905
1819
559
2537
3476
3395
1541
4485
4859
Eukaryotic
PLFA
pM/g
128
1
1
1
0.4
0.4
1
2
2
1
1
1
ND
5
7
ND
ND
2
2
1
3
9
8
6
10
ND
33
48
78
59
89
38
65
Prokaiyote/
Eukaryote
14
69
499
624
262
38
36
62
127
82
30
28
NO
768
262
NO
NO
356
269
453
344
416
454
666
384
NO
17
53
44
57
17
118
75
Gram+/
Anaerobic Gram-
(TerBrSats)
3.5
3.4
5.3
2.7
1.7
1.5
0.0
8.5
7.7
2.1
0.8
0.0
11.2
15.3
11.2
16.8
10.7
14.8
17.0
10.3
11.2
11.8
10.2
11.3
11.3
9.7
12.1
10.4
13.4
13.0
13.2
6.0
8.8
Community Structure (% of total PLFA)
Gram- Anaerobic SRB/ Genera
(Monos) Metal Reducers Actinomycetes Nsats
62.4
68.9
68.3
77.3
80.2
71.8
87.8
52.0
59.2
82.8
84.3
86.2
52.9
48.5
57.0
45.0
54.7
59.6
66.2
53.4
57.8
52.7
53.4
51.8
51.2
54.3
24.3
56.9
50.9
52.9
43.2
62.1
54.8
(BrMonos)
2.0
11.1
8.5
5.6
4.6
8.9
1.7
11.7
10.9
3.3
1.9
0.6
13.5
16.1
11.9
14.3
10.5
9.5
4.8
13.0
10.5
13.2
13.3
13.8
14.3
13.1
3.5
9.4
11.0
10.6
3.2
4.8
6.3
(MidBrSats)
2.1
3.4
3.0
1.9
1.3
2.7
0.0
6.7
6.7
1.8
1.2
0.0
7.0
6.7
5.3
7.0
4.8
5.3
1.6
6.6
5.2
4.7
5.3
4.9
4.7
4.9
26.9
4.0
4.7
4.1
15.6
1.8
2.7
23.2
12.0
14.8
12.3
11.9
12.7
7.8
19.5
14.8
8.7
8.6
9.7
15.4
13.2
14.2
16.9
19.4
10.5
10.1
16.4
15.1
17.5
17.7
18.1
18.3
18.1
23.0
17.5
17.9
17.6
19.4
24.4
26.1
Eukaryotes
(Polyenoics)
6.7
1.4
0.2
0.2
0.4
2.6
2.7
1.6
0.8
1.2
3.2
3.5
0.0
0.1
0.4
0.0
0.0
0.3
0.4
0.2
0.3
0.2
0.2
0.2
0.3
0.0
5.5
1.9
2.2
1.7
5.4
0.8
1.3
Physiological
Status
0.59
0.08
0.08
0.02
0.04
0.00
0.00
0.00
0.03
0.12
0.00
0.00
0.34
0.32
0.32
0.27
0.22
0.25
0.08
0.29
0.34
0.19
0.22
0.14
0.17
0.16
0.17
0.10
0.10
0.10
0.03
NO
0.10
The number of cells is calculated based on 2.0x1012 cells per gram dry weight of cells and 108 picolmoles of phospholipid per gram dry weight of cells. Microbial Insights (G. Davis, pers. comm.)
points out that the number of cells/gram of dry weight may vary and is dependent on the environmental conditions from which the microorganisms were recovered. Physiological status is based on ratios
of fatty acid biomarkers that indicate a metabolic response to environmentally induced stress evidenced by decreased membrane permeability (w7/w7c). The biomarker indicators generally suggest that the samples
have slow to moderate rates of turnover. Sample locations are shown in Figures 4.1 to 4.3. ND, not detected; NC not calculated.
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