&ER&
United States
Environmental Protection
Agency
       Monitored Natural Attenuation
       of Inorganic Contaminants in
       Ground Water
       Volume 2
       Assessment for Non-Radionuclides
       Including Arsenic, Cadmium, Chromium,
       Copper, Lead, Nickel, Nitrate,
       Perchlorate, and Selenium
        Methanogenic   Fe Reducing
          SO4 Reducing
             Mn Reducing
                          Denitrification
                                       Aerobic
        Precipitation
         of Sulfides
           Cr(VI) Reduction
          and Co-precipitation
           with Fe Oxides
 Nitrate
Attenuation
Adsorption to
Aquifer Solids

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                                              EPA/600/R-07/140
                                                October 2007
Monitored  Natural Attenuation of Inorganic
        Contaminants in Ground Water
                     Volume 2
 Assessment for Non-Radionuclides Including
Arsenic,  Cadmium, Chromium, Copper, Lead,
   Nickel, Nitrate,  Perchlorate, and  Selenium
                         Edited by
             Robert G. Ford, Richard T. Wilkin, & Robert W. Puls
                 U.S. Environmental Protection Agency
                 Office of Research and Development
              National Risk Management Research Laboratory
                     Ada, Oklahoma 74820
                        Project Officer
                        Robert G. Ford
             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, Ohio 45268

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                               Notice
    The U.S. Environmental Protection Agency through its Office of Research and
Development managed portions of the technical work described here under EPA
Contract No. 68-C-02-092 to Dynamac Corporation, Ada, Oklahoma through funds
provided by the U.S. Environmental Protection Agency's Office of Air and Radiation
and Office of Solid Waste and Emergency Response. It has been subjected to the
Agency's peer and administrative review and has been approved for publication
as an EPA document. Mention of trade names or commercial products does not
constitute endorsement or recommendation for use.

    All research projects  making conclusions  or recommendations based  on en-
vironmental data and funded by the U.S.  Environmental Protection Agency are
required to participate in the Agency Quality Assurance Program. This project did
not involve the collection or use of environmental data and, as such, did not require
a Quality Assurance Plan.

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                                             Foreword
The U.S. Environmental Protection Agency 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 environmen-
tal 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 is the Agency's center for investigation of technological and
management approaches for preventing and reducing risks from pollution that threatens human health and the
environment. The focus of the Laboratory's research program is on methods and their cost-effectiveness for pre-
vention 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.  NRMRLs 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 to assist the user community and to link researchers
with their clients. Understanding site characterization to support the use of monitored natural attenuation (MNA) for
remediating inorganic contaminants in ground water  is a major priority of research and technology transfer for the
U.S. Environmental Protection Agency's Office of Research and  Development and the National Risk Management
Research Laboratory. This document provides technical recommendations regarding the development of conceptual
site models and site characterization approaches useful for evaluating the effectiveness of the natural  attenuation
component of ground-water remedial actions.
                                           Stephen G. Schmelling, Director
                                           Ground Water and Ecosystems Restoration Division
                                           National Risk Management Research Laboratory

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                                     Contents
Foreword	iii
Figures 	vii
Tables 	ix
Acknowledgments 	x
Executive Summary 	xi


Chapter 1 - Cadmium 	1
    Occurrence and Distribution  	1
    Geochemistry and Attenuation Processes 	1
    Site Characterization	4
    Long-term Stability and Capacity	5
    Tiered Analysis  	6
    References	7
Chapter 2 - Lead 	11
    Occurrence and Distribution  	11
    Geochemistry and Attenuation Processes 	12
    Site Characterization	15
    Long-term Stability and Capacity	16
    Tiered Analysis  	17
    References	18
Chapters- Nickel	21
    Occurrence and Distribution  	21
    Geochemistry and Attenuation Processes 	22
    Site Characterization	25
    Long-term Stability and Capacity	27
    Tiered Analysis  	27
    References	28
Chapter 4 - Copper  	33
    Occurrence and Distribution  	33
    Geochemistry and Attenuation Processes 	34
    Site Characterization	36
    Long-term Stability and Capacity	37
    Tiered Analysis  	37
    References	38
Chapter 5 - Chromium	43
    Occurrence and Distribution  	43
    Geochemistry and Attenuation Processes 	43
    Site Characterization	47
    Long-term Stability and Capacity	50
    Tiered Analysis  	50
    References	51
Chapter 6 - Arsenic  	57
    Occurrence and Distribution  	57
    Geochemistry and Attenuation Processes 	58
    Site Characterization	61

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    Long-term Stability and Capacity	64
    Tiered Analysis  	64
    References	66
Chapter 7 - Selenium  	71
    Occurrence and Distribution 	71
    Geochemistry and Attenuation Processes 	73
    Site Characterization	75
    Long-term Stability and Capacity	79
    Tiered Analysis  	79
    References	80
Chapters- Nitrate	87
    Occurrence and Distribution 	87
    Geochemistry and Attenuation Processes 	91
    Site Characterization	92
    Long-term Capacity 	94
    Tiered Analysis  	95
    References	97
Chapter 9- Perchlorate 	101
    Occurrence and Distribution 	101
    Geochemistry and Attenuation Processes 	102
    Site Characterization	103
    Long-term Capacity 	105
    Tiered Analysis  	105
    References	106
                                           VI

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                                           Figures
Figure 1.1     Species distribution of Cd(ll) in pure water as a function of pH at 25 C	2

Figure 1.2     Cadmium speciation as a function of pH in solution containing chloride (100 mg L~1),
               sulfate (100 mg L~1), and inorganic carbon (100 mg L~1)	2

Figure 1.3     Eh-pH diagram for cadmium (total Cd = 10'5 molal, total C = 10'3 molal, total S = 10'3
              molal; all organic cadmium complexes are suppressed; activity coefficients for all
              species are set equal to 1)	3

Figure 2.1     Pb(ll) species distribution in pure water at 25 C	12

Figure 2.2     Species distribution of lead  in solution with 100 mg L1 chloride, 100 mg L1 sulfate,
              and 100 mg L1 total inorganic carbon, based on thermodynamic data in
              MINTEQA2 (Allison et al., 1990)	12

Figure 2.3.     Eh-pH diagram for lead (total Pb = 10-5molal, total  C = 10-3molal, total S = 10-3molal;
              all organic lead complexes are suppressed; activity coefficients for all species are
              set equal to 1)	13

Figure 2.4     Pb(ll) activity in equilibrium with PbCO3 (at total inorganic carbon equal to
              0.001 molal), PbSO4 (at total sulfate equal to 0.1 molal), and PbS (at total sulfide
              equal to 0.001  molal)	13

Figure 3.1     (a) Predicted solubility of various Ni precipitates that could form in aerobic ground
              water with concentrations of Al and Si controlled by the solubility of the clay
              mineral, kaolinite	23

Figure 3.2     Eh-pH diagrams for nickel at 25 C. (a) System Ni-H2O-Ca-AI-NO3-HCO3-SO4
              (2 mg Ni/L; 40 mg Ca/L; 3 mg AI/L; 6 mg NO./L; 60 mg HCO./L; 100 mg SO4/L)	23

Figure 3.3     Nickel sorption as a function of pH in the presence of an hypothetical aquifer
              sediment with iron  and manganese oxides reflective of the crustal abundance of
              these elements	24

Figure 4.1     Solubility of copper oxide and copper hydroxide as a function of pH in the system
              Cu-O-H at 25 C	34

Figure 4.2     Eh-pH diagram for copper at 25 C (total inorganic carbon = 10"2 molal; total
              sulfur = 10"3 molal; total copper = 10"5 molal)	35

Figure 4.3     Solubility and speciation of copper as a function of pH and log fugacity of CO2(gas)
              at25C	35

Figure 5.1     Distribution of 1  mM (millimoles per liter, which equals approximately 52 mg L1)
              chromium(VI) plotted as a function of pH in equilibrium with barite	44
                                               VII

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Figure 5.2     Calculated distribution of 1 mM Cr(VI) (approximately 52 mg L~1) in the presence
              of 1 mM K (approximately 39 mg L~1) and hydrous ferric oxide calculated as a
              function of pH	45

Figure 5.3     Concentration of dissolved Cr(VI) in equilibrium with Cr(VI) adsorbed on freshly
              precipitated hydrous ferric oxide	45

Figure 6.1     The distribution of arsenite and thioarsenic species in a reducing ground water	58

Figure 6.2     Eh-pH diagram for arsenic at 25 C	59

Figure 6.3     Eh-pH diagrams for arsenic and iron at 25 C for iron-reducing systems	60

Figure 6.4     Eh-pH diagrams for arsenic and iron at 25 C for coupled iron- and sulfate-reducing
              systems	60

Figure 7.1     Eh-pH diagram for selenium at 25 C using thermodynamic data from
              Seby  et al. (2001)	73

Figure 7.2     Eh-pH diagram for selenium at 25 C using thermodynamic data from
              Seby  et al. (2001)	74
                                               VIM

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                                           Tables



Table 1.1      Natural attenuation and mobilization pathways for cadmium	4

Table 2.1      Natural attenuation and mobilization pathways for lead	15

Table 3.1      Natural attenuation and mobilization pathways for nickel	25

Table 4.1      Natural attenuation and mobilization pathways for copper	37

Table 5.1      Natural attenuation and mobilization pathways for chromium	47

Table 5.2      Published USEPA methods for determination of total chromium and speciation in
              aqueous samples	48

Table 6.1      Natural attenuation and mobilization pathways for arsenic	62

Table 6.2      Examples of arsenic mobilization due to shifts in  ground-water chemistry	64

Table 7.1      Natural attenuation and mobilization pathways for selenium	76

Table 7.2      Published USEPA methods for determination of selenium in aqueous samples	76

Table 7.3      Review of selenium isotope fractionation ranges for abiotic-biotic processes
              during reduction and oxidation (Johnson, 2004)	78

Table 8.1      Example field applications of biotic remedial technologies for nitrate removal	89

Table 8.2      Natural attenuation and mobilization pathways for nitrate	90

Table 9.1      Natural attenuation and mobilization pathways for perchlorate	103

Table 9.2      Published USEPA methods for determination of perchlorate and other
              Cl-bearing inorganic ions in aqueous samples	104
                                               IX

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                                     Acknowledgments
This document represents a collective work of many individuals with expertise in the policy and technical aspects of
selecting and implementing cleanup remedies at sites with contaminated ground water. Preparation of the various
components of this document was undertaken by personnel from the USEPA Office of Research and Development
(ORD), Office of Superfund Remediation and Technology Innovation (OSRTI), and Office of Radiation and Indoor
Air (ORIA), as well as technical experts whose participation was supported under USEPA Contract No. 68-C-02-092
to Dynamac Corporation, Ada, Oklahoma through funds provided by ORIA and OSRTI. Contributing authors are
listed below along with contaminant chapters to which contributions were made:
  Contributing
     Author
                 Chapter

Richard T. Wilkin  Cadmium

                Lead
                Nickel
                Copper
                Arsenic

                Selenium

                Perchlorate

Robert G. Ford   Nickel

                Arsenic
                Selenium
                Nitrate
                Perchlorate
Chunming Su     Selenium

                Nitrate
                              Affiliation

USEPA/ORD, National Risk Management Research Laboratory, Ada, OK 74820
                           USEPA/ORD, National Risk Management Research Laboratory, Cincinnati, OH
                           45268
                           USEPA/ORD, National Risk Management Research Laboratory, Ada, OK 74820
                           USEPA/ORD, National Risk Management Research Laboratory, Cincinnati, OH
                           45268
Robert W. Puls   Chromium   USEPA/ORD, National Risk Management Research Laboratory, Ada, OK 74820

KirkG.Scheckel  Nickel

Douglas B. Kent  Lead       U.S. Geological Survey, McKelvay Building (MS-465), Menlo Park, CA 94025
                Chromium
                Arsenic
Patrick V. Brady   Lead

                Perchlorate
                           Sandia National Laboratories, Geochemistry Department (MS-0750),
                           Albuquerque, New Mexico 87185
Critical and constructive reviews were provided by Kenneth Lovelace (USEPA/OSRTI), John Washington (USEPA/
ORD National Exposure Research Laboratory, Athens, GA), Jackson  Ellington (USEPA/ORD National Exposure
Research Laboratory, Athens, GA), George Redden (Idaho National Laboratory), and Sue Clark (Washington State
University). Pat Bush (Ada, OK) is acknowledged for her technical editing to provide consistency in formatting and
grammar.  Martha Williams (Contract #68-W-01-032) assisted with final editing and formatting for publication.

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                                      Executive Summary
The term "monitored natural attenuation," as used in this document and in the Office of Solid Waste and Emergency
Response (OSWER) Directive 9200.4-17P, refers to "the reliance on natural attenuation processes  (within the con-
text of a carefully controlled and monitored site cleanup approach) to achieve site-specific remediation objectives
within a time frame that is reasonable compared to that offered by other more active methods." When properly em-
ployed, monitored natural attenuation (MNA) may provide an effective knowledge-based remedy where a thorough
engineering analysis informs the understanding, monitoring, predicting, and documenting of the natural processes.
In order to properly employ this remedy, the Environmental Protection Agency needs a strong scientific basis sup-
ported by appropriate research and site-specific monitoring implemented in accordance with the Agency's Quality
System.  The purpose of this series of documents, collectively titled "Monitored Natural Attenuation of Inorganic
Contaminants in Ground Water," is to provide a technical resource for remedial site managers to define and assess
the potential for use of site-specific  natural processes to play a role in the design of an overall remedial approach
to achieve cleanup objectives.

The current document represents the second volume of a set of three volumes that address the technical basis and
requirements for assessing the potential applicability of MNA as part of a ground-water remedy for plumes with non-
radionuclide and/or radionuclide inorganic contaminants. Volume 2, titled "Assessment for Non-Radionuclides Includ-
ing Arsenic, Cadmium, Chromium, Copper, Lead, Nickel, Nitrate, Perchlorate, and Selenium," consists of individual
chapters  that describe 1) the natural processes that may result in the attenuation of the listed contaminants and 2)
data requirements to be met during site characterization. Emphasis is placed on characterization of immobilization
and/or degradation processes that may control contaminant attenuation, as well as technical approaches to assess
performance characteristics of the MNA remedy. A tiered analysis approach  is presented to assist  in organizing
site characterization tasks in a manner designed to reduce uncertainty in remedy selection while distributing  costs
to address four primary issues:

    1.  Demonstration of active contaminant removal from ground water & dissolved plume stability;
    2.  Determination of the rate and  mechanism of attenuation;
    3.  Determination of the long-term capacity for attenuation  and stability of immobilized contaminants;
       and
    4.  Design of performance monitoring program, including defining triggers for assessing  MNA  failure, and
       establishing a contingency plan.
Where feasible, Agency-approved  analytical protocols currently implemented for waste site characterization are
identified, along with modifications that may be warranted to help insure the quality of site-specific data.  In situ-
ations where Agency methods  or protocols are unavailable, recommendations are  made based on  review of the
existing technical literature. It is anticipated that future updates to these  recommendations may be warranted with
increased experience in the successful application of MNA as part of a ground-water remedy and the development
of new analytical protocols.

This document is  limited to evaluations performed  in porous-media settings. Detailed discussion  of performance
monitoring system design  in  fractured rock, karst,  and other such highly heterogeneous settings is beyond the
scope of this document. Ground water and contaminants often move preferentially through discrete  pathways (e.g.,
solution channels, fractures, and joints) in these settings.  Existing techniques may  be incapable of fully delineat-
ing the pathways  along  which  contaminated  ground water migrates. This greatly  increases the uncertainty and
costs of assessments of contaminant  migration and fate and is another area  of continuing research. As noted in
OSWER  Directive 9200.4-17P, "MNA will  not generally be appropriate where site complexities preclude adequate
monitoring." The directive provides additional  discussion regarding the types of sites where the use of MNA may
be appropriate.

This document focuses on  monitoring  the saturated zone,  but site characterization and monitoring  for MNA or any
other remedy typically would include monitoring of all significant pathways by which contaminants  may move from
source areas and contaminant plumes to impact receptors (e.g., surface water and indoor air).
                                                  XI

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Nothing in this document changes Agency policy regarding remedial selection criteria, remedial expectations, or the
selection and implementation of MNA.This document does not supercede any guidance. It is intended for use as a
technical reference in conjunction with other documents, including OSWER Directive 9200.4-17P, "Use of Monitored
Natural Attenuation at Superfund, RCRA Corrective Action, and Underground Storage Tank Sites" (http://www.epa.
gov/swerust1/directiv/d9200417.pdf).
                                                  XII

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                                              Cadmium
                                          Richard T. Wilkin
Occurrence and Distribution

Cadmium is comparatively rare in the environment with an
average abundance similar to other second- and third-row
transition metals (e.g., silver and mercury). The  median
concentration of cadmium in soils and sediments ranges
from about 0.04 to  1.8  mg kg-1 (Reimann and  Caritat,
1998).  Where cadmium concentrations are elevated, it is
typically found in association with sulfide ores of zinc and
sometimes with ores of  copper and lead.  The primary
mineral associations of cadmium are with otavite (CdCO3),
greenockite (CdS), sphalerite (ZnS), smithsonite (ZnCO3),
and hemimorphite (Zn4Si2O7(OH)2-H2O). Soil weathering
can lead to  release of the Cd2+ ion, which is generally
soluble and mobile in water.

The primary  industrial uses of cadmium are metal  plating,
production of Ni-Cd batteries, as a stabilizer in plastics, and
as a pigment. According  to Minerals Information statistics
for 2001, approximately  75% of the U.S. apparent con-
sumption of cadmium (a total of about 2.4 million pounds)
went into production of  Ni-Cd batteries (Wolke, 2003).
The largest sources  of cadmium contamination to ground
water and surface water  are from sewage sludge, mines
(e.g., mine water, mine tailings leachate), metal smelters
(process waters), battery recycling plants, and wastes from
electroplating facilities.

Remedial Technologies

Treatment of cadmium and other heavy metals in industrial
wastewater streams is often achieved by precipitation using
lime, sodium carbonate,  alkaline sulfides, or  organosul-
fides.  These treatment methods are generally unsuitable
for drinking water.  Ion exchange resins and  adsorption
substrates are in most cases used  for treatment of drink-
ing water contaminated with  cadmium (Zhao et al., 2002;
Lai et al., 2002). Technology classes suitable for remedia-
tion of cadmium-contaminated soils include containment,
solidification/stabilization, and  separation/concentration
(USEPA, 1997). Containment technologies applied at metal
contamination sites to minimize the transport of cadmium
and co-contaminants out of source zones include caps
and vertical barriers. Reactive barriers are appropriate for
treatment of  some cadmium and co-contaminant  ground
water plumes (e.g., Gibertetal., 2003; Wang and Reardon,
2001).
Regulatory Aspects

Cadmium and its compounds are very toxic to nearly all liv-
ing organisms. The EPA has set the maximum contaminant
level (MCL) for cadmium in drinking water at 0.005 mg L1.
Cadmium  is fairly mobile  and soluble in water at low to
near-neutral pH.  The main  routes by  which cadmium
enters the human body are ingestion of plant-based food
and inhalation of cadmium-bearing dusts.  The kidney is
the primary organ affected by exposure to cadmium.  For
non-potable water sources, ambient water quality criteria
(AWQC) that are protective of aquatic life  may serve as
alternative cleanup goals.  For cadmium, current statutes
list both acute and  chronic criteria for fresh waters as
0.002 mg L1 and 0.00025  mg L1, respectively, for a water
hardness of 100 mg L1  (USEPA, 2006; http://www.epa.
gov/waterscience/criteria/nrwqc-2006.pdf).   Adjustments
to these criteria are to be applied for waters with different
hardness.  An example of where this criterion may apply
is a site where contaminated  ground water discharges to
surface water.

Geochemistry and Attenuation Processes

Aqueous Spec/at/on

Dissolved  forms of cadmium  are  only present in the +2
valence state.  Cadmium  has a tendency to form aque-
ous complexes with both  inorganic and organic ligands,
although the uncomplexed Cd2+ ion is fairly stable.  The
most important inorganic cadmium complexes are with hy-
droxide, chloride, sulfate, bicarbonate, carbonate, cyanide,
and ammonia. Complexation of cadmium with humic acids
is important under conditions of  high dissolved organic
carbon (DOC) concentrations, but binding of cadmium  with
humic acids appears to be weaker when compared to  lead
(Abate and Masini, 2002; Christensen and Christensen,
1999; Dunnivant et al., 1992).  In highly reducing systems,
cadmium complexation with bisulfide is possible. It is likely
that cadmium toxicity is related to its strong tendency to
form bonds with thiol functional groups in certain enzymes
which results in the displacement of biologically essential
metals (Baes and  Mesmer, 1976).

The fractional abundance of Cd-OH species in water as a
function of pH is shown in Figure 1.1. The distribution dia-
gram for cadmium hydroxy complexes indicates that Cd2+,
Cd(OH)+, Cd(OH)2 are the most significant species below

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pH 12. At low total concentrations of cadmium, hydrolysis of
Cd2+ becomes significant above about pH 9. Complexation
of cadmium with chloride and ammonia becomes important
as aqueous concentrations of these ligands exceed 10~2
molal and 10'3 molal, respectively  (e.g., Lindsay, 1979).
   1.fl-
.a
CO
g
13
                           PH
Figure 1.1  Species distribution of Cd(ll) in pure water
            as a function of pH at 25 C.


Figure 1.2 shows the  fractional  abundance of cadmium
species as a function of pH in an aqueous solution con-
taining a  mixture of chloride, sulfate, and  inorganic car-
bon.  Again  the uncomplexed Cd2+ ion dominates below
pH 8.5. In general, complexation  of cadmium with chloride
and sulfate is most  important at near-neutral to low pH;
carbonate complexation is most  important  at pH 9 to 11;
and, hydroxyl species dominate cadmium complexation  at
pH>11 (Figure 1.2).

The identity of cadmium bisulfide complexes and their
formation constants  have been discussed  by  Daskalakis
and Helz (1992) and Wang and Tessier (1999).  Uncer-
tainty  persists  regarding  the stoichiometry of the  most
important cadmium  complexes  in sulfidic waters.  This
uncertainty mainly stems from the experimental approach
that has been traditionally used to extract thermodynamic
data, i.e., evaluation  of CdS solubility over a range of total
cadmium concentrations, total sulfide concentrations, and
pH. Data presented in Wang and Tessier (1999) indicate
that Cd(HS)2 is the dominant species at ES(-II)=10-5 molal
and over the pH range from 6 to 8, typically encountered in
natural sulfidic waters. At lower total sulfide concentrations,
CdHS+ and Cd2+ become increasingly important (Wang and
Tessier, 1999).
                                                      CO
                                                      a
                                                      c
                                                      1
                                                      CO
                                                      O
                                                      13
   0.8-
   0.6-
                                                         0.4-
                                                         0.2-
                                                         0.0-
                                                             Cd-CL
                                                                     Cd-S04
                                                                                              10
                                                                                                       12
                                                                                 PH
Figure 1 .2  Cadmium speciation as a function of pH in
            solution containing chloride (100 mg L1),
            sulfate (100 mg L1), and inorganic carbon
            (100 mg L1). Cadmium chloride complexes
            include CdCI+ and CdCI.  Cadmium sulfate
            complexes include CdSO and CdfSOJ/:
            Cadmium complexes with inorganic carbon
            include CdCO, CdHCO+, and CdfCOJ/:
            Cadmium hydroxy complexes include
            CdOH+, Cd(OH)2, and Cd(OH)3. Total
            cadmium is equal to 1 mg L1.

Solubility

An Eh-pH diagram for cadmium is shown in  Figure 1.3.
Inspection of this diagram indicates that at the specified
conditions Cd2+ is the soluble form of cadmium at pH < 5
and at moderate to highly oxidizing redox potentials. At
near-neutral to moderately alkaline pH (6 to about  12)
cadmium carbonate (otavite) is stable, and at pH >  12.5
cadmium hydroxide is stable.  In sulfidic environments,
cadmium sulfide  (greenockite) is stable over a  wide pH
range. Solubility  expressions for cadmium carbonate and
cadmium sulfide are given by:
CdCO
                     = Cd2+ + HCO3-     (log K= -0.9)
and
            CdS + H+ = Cd2+ + HS-      (log K= -14.4)

In natural deposits, cadmium often substitutes for zinc in
the mineral structures of sphalerite (ZnS) and smithsonite
(ZnCO3). O'Day et al. (1998) suggest that as cadmium-
substituted  sphalerite weathers, cadmium is preferentially
partitioned  into the aqueous phase over zinc. Zinc was
found to form various  zinc hydroxides  and/or  zinc-iron
oxyhydroxides depending on the total amount of iron in
the system. Cadmium was not identified in the solid-phase
products from weathering indicating its general tendency
to be mobile in the aqueous phase (O'Day et al., 1998;
Carroll etal., 1998).

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  LU
     -.5
                                    10
                                          12
                                                14
                           PH
Figure 1.3  Eh-pH diagram for cadmium (total Cd =
            10-5 molal, total C = 10-3molal, total S = W3
            mo/a/; all organic cadmium complexes are
            suppressed; activity coefficients for all spe-
            cies are set equal to  1).

Cadmium carbonates - Cadmium is known to form solid-so-
lutions with calcium carbonate (calcite). The Cd2+ and Ca2+
ions are nearly the same size with crystal radii of 1.09 and
1.14 angstroms, respectively, assuming octahedral coordi-
nation. The formation of Ca-Cd carbonate solid solutions
is environmentally significant because solid solutions are
generally more stable and  less soluble than endmember
compositions.  Cadmium uptake from aqueous solution by
calcite has been widely studied (e.g., McBride, 1980; Davis
et al., 1987; Papadopoulos and Rowell, 1988;  Zachara et
al.,  1991; Stipp et al., 1992, 1993; Tesoriero and Pankow,
1996; Chiarello et al.,  1997; Martin-Garin et  al.,  2003).
Cadmium uptake is  thought to consist of two  processes.
The first process is rapid sorption and the second process
is incorporation into the crystal lattice and formation of an
otavite-calcite solid-solution. The latter process has been
confirmed through determination of cadmium solid phase
speciation during reaction with calcite (Bailey et al., 2005).
Cadmium partitioning to carbonaceous materials has been
applied by Wang and Reardon (2001) for the  removal of
cadmium from wastewater streams.

Cadmium phosphates  - Santillan-Medrano and Jurinak
(1975) observed the formation  of  cadmium  phosphate
precipitates in soil systems containing phosphorus. Pre-
cipitation of phosphate  compounds  over carbonate com-
pounds was favored in phosphorous-containing systems at
pH < 7.5. Substitution of cadmium into natural  apatite has
also been documented  (Sery et al., 1996) and may be a
more common route for partitioning to phosphate minerals
at concentrations undersaturated with respect to precipita-
tion of cadmium phosphate.
Cadmium hydroxides- Baes and Mesmer (1976) report the
existence of three crystalline forms of Cd(OH)2.  At 25 C
the stable form is p-Cd(OH)2. This material is fairly soluble
at circumneutral pH.  Cadmium concentrations below the
MCL of 0.005 mg L1 would only be approached at pH > 11
in systems saturated with respect to |3-Cd(OH)2.

Cadmium sulfides - It is well known that Cd2+ and HS~ react
to form a very insoluble, yellow-colored precipitate.  How-
ever, Daskalakis and Helz (1992) point out that under some
conditions dissolved cadmium bisulfide complexes are very
stable; consequently, the effectiveness of sulfide treatment
for cadmium in wastewater has been overestimated in some
cases due to uncertainty regarding the stability of cadmium
bisulfide species.

Framson and Leckie (1978) examined the limit of coprecipi-
tation of cadmium and ferrous monosulfide. Their experi-
mental data suggest only limited coprecipitation,  likely due
to the size mismatch between the ferrous (~0.75 angstroms)
and cadmium (~1.09 angstroms) cations. They suggest that
in sulfidic systems, cadmium precipitates primarily through
surface exchange with ferrous monosulfide substrates or
as unsubstituted cadmium sulfide. Parkman et al. (1999)
performed X-ray absorption  spectroscopy experiments
and also concluded that  a CdS phase  is formed as  cad-
mium interacts with iron monosulfide. On the other hand,
Coles et  al. (2000) found up to 29% replacement of iron
by cadmium in mackinawite (FeS). It is possible that this
high percentage of replacement occurs at the surface of
very fine-grained iron sulfide particles. Coles et al. (2000)
found that the mixed ferrous-cadmium  sulfides are more
insoluble than pure mackinawite.

Adsorption

Adsorption/desorption behavior of cadmium  is strongly
a function of pH, and to a lesser extent a function of the
solution concentration of cadmium and the concentration
of competing cations or complexing ligands. At low  con-
centrations  of cadmium,  sharp adsorption edges provide
evidence that cadmium forms strong bonds with mineral
surfaces. In general, the presence of calcium and mag-
nesium reduces the extent of cadmium removal by aquifer
solids. The presence or addition of zinc, which  tends to be
more strongly adsorbed, can reduce the amount  of cadmium
uptake by iron and aluminum oxides, indicating that zinc
competes for similar adsorption sites and is preferentially
adsorbed over cadmium  (Benjamin and Leckie,  1980). It
has also  been observed  that  cadmium  may preferentially
adsorb to manganese oxides when  they  are present in
sufficient quantities (Bellanca et al.,  1996). Tonkin et al.
(2004) have evaluated published cadmium adsorption data
to determine surface complexation constants that may be
employed to assess the potential extent  of adsorption onto
manganese oxides.

Ainsworth et al. (1994) examined the sorption  behavior of
cadmium on freshly prepared and aged hydrous ferric oxide
(HFO). Their results indicate that HFO effectively removes
cadmium from solution at pH above about 6.7.  In general,
the pH-dependent adsorption behavior parallels the change

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in  aqueous speciation from Cd2+ to cadmium hydroxide
species (see Fig. 1.1), although adsorption occurs at pH
values where cadmium hydroxide species are unexpected
in bulk solution. Aging times of up to 21 weeks showed little
effect on the sorption behavior of cadmium onto HFO. HFO
aged in the presence of Cd2+ ions showed some desorption
hysteresis suggesting that cadmium is incorporated into the
metal oxide structure during recrystallization. Martfnez and
McBride (1998) suggest that coprecipitation of cadmium
with amorphous iron oxides results in more reduced con-
centrations than can be achieved through surface adsorp-
tion alone. However, Ford  et al. (1997) report that during
long-term aging of hydrous iron oxides, cadmium desorbs
or is released suggesting minimal incorporation of cadmium
into the goethite or hematite structures.

Lai et al. (2002) investigated the adsorption characteristics
of cadmium and humic acid onto iron oxyhydroxide-coated
quartz sands. The adsorption of both cadmium and humic
acid was highly pH dependent.  Cadmium adsorption  in-
creased with pH, whereas humic acid adsorption decreased
as pH increased.  The presence of humic acid was found
to  result in increasing cadmium adsorption capacity in the
pH range of 4-6.

Redox Chemistry

In  natural systems cadmium is present in the +2 oxidation
state. Therefore, the geochemical transport processes of
cadmium are not  directly tied to changes in  redox condi-
tions.  Because  cadmium  forms stable precipitates and
aqueous complexes with redox-sensitive elements such as
sulfur and carbon, its mobility potential is indirectly tied to
redox conditions.  In sulfate-reducing systems, cadmium is
expected to form  insoluble CdS precipitates  or coprecipi-
tates with FeS (DiToro et al., 1990). In moderately reducing
                           but non-sulfidic systems, however, reductive dissolution of
                           hydrous ferric oxides with adsorbed cadmium could result
                           in cadmium mobilization.

                           Several studies indicate that concentrations of dissolved
                           cadmium increase when reduced systems are  oxidized,
                           such as when dredged sediments are land filled. This be-
                           havior may be due to oxidative dissolution of metal sulfides
                           or due to the decomposition of organic materials that bind
                           cadmium (e.g., Cooper and Morse, 1998; Simpson et al.,
                           2000; Martinez et al., 2002).

                           Colloidal Transport

                           Transport of cadmium via  colloids can be significant in
                           ground water and surface water systems. Both mineral and
                           organic particles can play a  role in binding and transporting
                           cadmium.  Cadmium  adsorbed to colloidal hydrous ferric
                           oxides may subsequently desorb due to pH decreases or
                           due to decreases in the oxidation-reduction potential.

                           Site Characterization

                           Cadmium mobility in ground water is governed by the total
                           concentration of cadmium, the distribution of cadmium spe-
                           cies in water, and the nature of cadmium partitioning in the
                           solid phase. The development of site-conceptual models
                           for predicting the long-term fate of cadmium at a contami-
                           nated site will require information on the  distribution and
                           concentration of cadmium in the aqueous phase and the
                           solid phase. Table 1.1 indicates possible  natural attenua-
                           tion  and mobilization pathways for cadmium. Details of the
                           types of analytical measurements that may be conducted
                           on sampled ground water and aquifer sediments to assist
                           in identifying the attenuation mechanism(s) are discussed
                           in the following paragraphs.
Table 1.1  Natural Attenuation and Mobilization Pathways for Cadmium
  Attenuation Processes
      Mobilization Processes
        Characterization Approach
 Precipitation of  insoluble
 carbonates, sulfides, and
 hydroxides. In general, pH >
 8 will drive precipitation reac-
 tions resulting in Cd concen-
 trations below the MCL.
Dissolution of carbonates at  low pH;
oxidative dissolution of sulfides at low to
neutral pH and high Eh; degradation of
organic matter; complexation/stabiliza-
tion in the presence of DOC.
Evaluation  of cadmium  concentration in  the
aqueous phase. Determination of total Cd in the
solid matrix. Evaluation of solid phase partition-
ing using sequential extraction methodologies.
Evaluation of long-term storage capacity.
 Sorption (reversible) to iron
 hydroxides, organic matter,
 carbonates, sulfides (pH > 6
 to 8).  Substitution for Zn in
 minerals.
Desorption at low pH; complexation/sta-
bilization in the presence of DOC. Re-
ductive dissolution of iron hydroxides.
Oxidation of metal sulfides.
Evaluation of cadmium speciation in the aque-
ous phase. Determination of total Cd in the solid
matrix.  Evaluation of solid phase partitioning
using sequential extraction  methodologies.
Batch and column testing to determine Cd up-
take capacity of site-specific aquifer  materials
at variable geochemical conditions.

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Aqueous Measurements
Quantitative measurement of cadmium concentrations  in
aqueous solutions is typically carried out using inductively
coupled plasma atomic emission spectroscopy (ICP-AES),
inductively coupled plasma mass spectroscopy (ICP-MS),
or atomic absorption spectroscopy (AAS).  Input data to
geochemical codes (e.g., MINTEQA2, PHREEQC, EQ3/6)
for  determining aqueous speciation also  require, at  a
minimum, concentrations of major anions, major cations,
dissolved organic carbon, temperature, and pH.  In addi-
tion, while regulatory requirements stipulate that unfiltered
ground-water samples be analyzed to support regulatory
decisions at a contaminated site, it may be necessary to
also collect filtered samples to help define the process(es)
controlling contaminant mobility. The use of 0.45  urn pore
size filters is  common as an arbitrary cutoff point  to dif-
ferentiate between dissolved and particulate  phases  in
water samples. However, caution is recommended when
using this approach, particularly for Fe and Al and other
elements that may be associated  with Fe or Al particles
(including Cd) that could pass through 0.45 um filters. The
use of filters with pore sizes less than 0.1 um will generally
provide a better assessment of the dissolved vs. particulate
load in ground water.

Solid Phase Measurements
The implementation of an analytical approach to identify
cadmium speciation in aquifer sediments is a challenging
process. The accuracy of the analytical finding is depen-
dent  on the  method  of  sample  collection/preservation
and the tools used to identify the mechanism of cadmium
partitioning. It is recommended that the analytical protocol
be designed  to address the potential redox sensitivity of
the solid phase(s) to which cadmium may be  partitioned
(e.g.,  sulfides in reduced sediments). Tools to evaluate the
mechanism of cadmium solid phase partitioning  range  in
complexity  from relatively simple chemical extractions to
advanced spectroscopic techniques.

The total concentration of cadmium in soils, sediments, and
aquifer materials may be determined by X-ray fluorescence
spectroscopy (XRF), or  by ICP-AES after digestion  in
mineral acids. A variety of digestion or extraction methods
can be found in the literature (Amacher,  1996).  Neutron
activation analysis is not commonly employed  due to the
scarcity of neutron sources required to irradiate the sample.
X-ray fluorescence is the most attractive approach due
to the relative ease of sample preparation, which may be
conducted with the sample in its original state or following
fusion with  lithium metaborate.   When combined  with the
determination of other major or  trace elements  in the solid
sample, this provides an initial step for assessing  possible
association of cadmium with various solid phase compo-
nents. This type of analysis can be conducted on the bulk
sample as well as at a microscopic level using wavelength
(electron microprobe)  or energy dispersive spectroscopy
coupled to a scanning or transmission electron microscope.
Microscopic examination allows one to better differentiate
whether cadmium may be distributed across a number of
different mineral phases within the solid sample or primar-
ily associated with a discrete phase.  There are limitations
to this approach (Pye, 2004), a significant one being that
the analysis does not necessarily provide unique mineral
identification  necessitating  the  collection of supporting
mineralogical and chemical data.

More detailed  information  on  the  specific  partitioning
mechanism(s) controlling cadmium solid phase speciation
is typically required to adequately support  site assess-
ment for potential reliance on natural attenuation as part
of a site remedy. There have been many applications of
sequential extraction schemes to  assess the speciation
of solid phase cadmium (e.g., Tessier et al., 1979; Mickey
and  Kittrick, 1984; Pustisek et al.,  2001; Buanuam et al.,
2006). As discussed in the cited reports, sequential extrac-
tion methods provide a  useful tool to assist in determining
the chemical speciation of trace metals in soils/sediments,
but essentially all documented methods show  analytical
limitations in selectively  extracting  cadmium  and  other
metals associated with  specific solid components.  Where
feasible, it is recommended that complimentary analytical
techniques be employed to  confirm the accuracy of cad-
mium speciation (e.g.,  O'Day et al.,  1998; O'Day  et al.,
2000; Carroll et al., 2002; D'Amore et  al., 2005) or the
accuracy of the extraction of a targeted phase(es) for a
given extractant (e.g.,  Shannon and White, 1991; Ngiam
and Lim, 2001; Peltier et al., 2005). As an example, Peltier
et al. (2005) have demonstrated that a common extraction
method  employed to target metals associated with easily
reducible iron (hydr)oxides may also dissolve iron sulfides
that  may be present. The results from this analysis may
lead to misidentification of a cadmium association with iron
(hydr)oxides, resulting in the development of a conceptual
site model that misrepresents the site-specific attenuation
process. Under reducing  conditions, it is also critical that
aquifer sediments be sampled and processed in a manner
that prevents exposure to oxygen prior to extraction in order
to limit oxidation  of reduced minerals (e.g., iron sulfides) that
may host cadmium. Cadmium associated with a sulfidic
phase in sediments has been shown to repartition to more
extractable phases  upon  oxidation (Saeki et al., 1993).
Determination of the  host  mineral  phase(es)  dissolved
for each extraction step is recommended, along with the
use of surrogate Cd-bearing phases spiked into the sedi-
ment to  confirm accuracy of the procedure (e.g., Rudd et
al., 1988). The  choice of  appropriate cadmium surrogate
phases  would be governed by site-specific geochemical
conditions or characterization of the mineralogy of the
aquifer sediment.

Long-Term Stability and Capacity

The stability of attenuated cadmium will largely depend on
the stability of site-specific geochemical conditions through
time. For  example,  if cadmium attenuation follows a cal-
cium carbonate coprecipitation  pathway,  then  long-term
stability  of attenuated  cadmium will depend, in part, on
the persistence  of pH conditions. If pH conditions were to
shift significantly to low values, cadmium might be expected
to release from  the solid phase. It  is therefore  important
to understand the attenuation mechanism(s) so that geo-

-------
chemical triggers for mobilization can be anticipated and
incorporated into evaluations of long-term monitoring data.
For any proposed and identified attenuation mechanism,
there will exist possible scenarios whereby remobilization
could occur (i.e., changes in pH or Eh). It will be essential to
explore the likelihood of such changes in site geochemistry
and the sensitivity of the attenuation pathway to changes
in the prevailing geochemical conditions.
Quantifying the attenuation capacity (as defined in Volume
1) will also require an understanding of the specific attenu-
ation pathway(s). Attenuation capacity, for example, could
be related to the extent that pH is buffered, the availability
of sorptive sites in aquifer materials, or to the supply of
electron donors needed to sustain microbially mediated re-
dox conditions.  For any proposed attenuation mechanism,
there will be assumptions built into capacity estimations,
so it is recommended that uncertainty analysis accompany
capacity calculations.

Tiered Analysis
Determination of the viability of cadmium remediation via
monitored natural attenuation will depend upon proper as-
sessment of contaminant loading to the aquifer and prevail-
ing  geochemistry and mineralogy within the contaminant
plume and the down gradient zone prior to the point(s) of
compliance. MNA may not  be  appropriate as a site rem-
edy for cadmium contamination in acidic to circum-neutral
pH, highly oxidizing, and/or DOC-rich environments.  The
goal of site assessment is to demonstrate the process(es)
controlling cadmium sequestration onto aquifer solids and
the long-term stability of solid phase cadmium as a function
of existing and anticipated ground-water chemistry.  The
following tiered analysis structure for site characterization
provides an approach to evaluate candidate sites and de-
fine the potential limitations of  MNA as part of a remedy
for ground-water cleanup.
Tier I - Site characterization under Tier I will involve dem-
onstration that the plume is static or  shrinking, has  not
reached  compliance boundaries,  and does not impact
existing water supplies.  Once this is established through
ground-water characterization, evidence  is collected to
demonstrate Cd partitioning to aquifer solids within  the
plume.  Rapid movement of contaminants along preferred
flow paths in the unsaturated and saturated zones can by
induced by hydrologic events such as heavy rains. It will be
important to determine that such hydrogeologic features do
not  result in contaminants bypassing zones where natural
attenuation is occurring. If natural attenuation processes
are active throughout the plume, then there should be an
observed increase in solid phase concentrations within
regions of the plume with higher aqueous concentrations,
e.g., near the source term. This field partitioning data
may be supplemented by geochemical modeling that in-
corporates measured water chemistry (e.g., pH, Eh, and
major ion chemistry) throughout the  plume to assess the
potential for solubility control by a cadmium precipitate such
as a carbonate/phosphate or sulfide  phase. Identification
of active sequestration to prevent cadmium migration in
ground-water provides justification for proceeding to Tier
II characterization efforts.

Tier II - Under Tier II, the apparent rate and mechanism(s)
of attenuation are determined.  Estimates of a site attenu-
ation rate(s) can  be assessed via a well transect along
the ground-water  flow path.  In addition, time-series data
may be collected  at one or more monitoring points within
the plume. This information  will allow assessment of the
relative timescales for contaminant immobilization and fluid
transport and determination of whether remediation objec-
tives can be met within the required regulatory time frame.
In addition, the mechanism(s) for attenuation need to be
identified under this stage of site characterization.  This
effort may require  determination of the chemical speciation
of aqueous and solid phase Cd, which may be approached
according to the following scheme:

 1)  Determination of cadmium solution speciation via
    direct analytical measurements in combination with
    speciation  calculations  based  on  characterized
    ground-water chemistry;
 2)  Calculation of the saturation state of ground water
    relative to measured aqueous chemistry compli-
    mented by the possible isolation of  discrete Cd
    mineral phases via density separations (or other
    schemes) in  regions of the aquifer with highest
    solid phase concentrations;
 3)  Determination of aquifer mineralogy to determine
    the relative abundance of components with docu-
    mented capacity for Cd sorption (e.g., Amonette,
    2002);
 4)  Identification  of cadmium  association(s) with the
    various solid  phase components of aquifer solids
    through combination of  chemical extractions with
    microscopic/spectroscopic confirmation of  phase
    associations, and;
 5)  Demonstration  of concurrence  between the site
    conceptual model and mathematical model(s) that
    describe cadmium removal mechanism(s).
It is recommended that identification of cadmium chemical
speciation in aqueous and  solid matrices be conducted
using  samples collected in a manner that preserves the
in-situ distribution of dissolved cadmium and mineralogy
and prevents loss of cadmium  from aqueous samples
(e.g., due to oxidation and precipitation of ferrous iron in
anoxic ground water).  The demonstration  of concurrence
between  conceptual and mathematical models describing
cadmium transport will entail development of site-specific
parameterization  of the chemical processes controlling
cadmium solid phase partitioning.

Tier III -  Once the partitioning mechanism(s) have been
identified for the site, the subsequent characterization effort
under Tier III will  involve determination  of the stability of
immobilized Cd and the capacity of the  aquifer  to sustain
continued uptake. It is recommended that the stability of
immobilized Cd be tested based on the anticipated evolution
of ground-water chemistry concurrent with plume shrinkage.

-------
For example, changes in ground-water pH can exert a sig-
nificant influence on Cd adsorption or precipitate solubility.
Therefore, it is recommended that sediment leach tests be
conducted to characterize the magnitude of Cd mobilization
as a function of pH for a ground-water chemistry representa-
tive of site conditions. It is recommended that the capacity
for Cd uptake onto aquifer solids be determined relative to
the specific mechanism(s) identified in Tier II. For example,
if site characterization under Tier II indicated that precipita-
tion of Cd sulfide due to microbial degradation of organic
compounds coupled with sulfate reduction occurs within the
aquifer, then it is recommended that the mass distribution
of organic carbon and sulfate to support this reaction within
the aquifer be determined. This site-specific capacity can
then be compared to Cd mass loading within the  plume
in order to assess the longevity of the natural attenuation
process. If site-specific  tests demonstrate the stability of
immobilized Cd and sufficient capacity within the aquifer to
sustain Cd attenuation, then the site characterization effort
can progress to Tier IV.  For cases where contaminant stabil-
ity is sufficient but aquifer capacity is insufficient for capture
of the entire plume, then a determination of the benefits of
contaminant source reduction may be necessary.

Tier IV- Finally,  under Tier IV a monitoring plan is  estab-
lished along with contingency plans  in the  event of MNA
failure.  It is recommended that  the monitoring plan be
designed to establish both continued plume stability and to
identify changes  in ground-water chemistry that may lead
to re-mobilization of attenuated Cd. The specific chemical
parameters to be monitored will include those identified un-
der Tier III that may halt Cd partitioning to aquifer sediments
and/or result in solubilization of either discrete Cd precipi-
tates or aquifer minerals that sequester Cd from ground
water. For example, solution phase parameters that could
alter either Cd precipitation or adsorption include increases
in soluble organic carbon in combination with changes in
ground-water pH. In contrast, the concentration of dissolved
iron or sulfate may indicate the dissolution of an important
sorptive phase within the aquifer (e.g., reductive dissolution
of iron oxides or oxidative dissolution of su If ides). Changes
in these parameters may occur prior to observed changes
in solution Cd and, thus, serve as monitoring triggers for
potential MNA failure.  In this instance, a contingency plan
can be implemented that incorporates strategies to arrest
possible plume expansion beyond compliance boundaries.
Possible strategies  to prevent plume  expansion include
pump and treat operations, installation of reactive barriers
to enhance uptake capacity perpendicular to the direction
of plume advance, or enhancement of natural attenuation
processes within  the aquifer through the injection of soluble
reactive components.

References

Abate,  G. and J.C.  Masini.  Complexation of  Cd(ll)  and
    Pb(ll) with humic  acids studied by  anodic stripping
    voltametry using differential equilibrium functions and
    discrete models. Organic Geochemistry 33:1171 -1182
    (2002).
Ainsworth, C.C., J.L. Pilon, PLGassman.andW.G.Vander
    Sluys. Cobalt, cadmium, and lead sorption to hydrous
    iron oxide: residence time effect. So/7 Science Society
    of America 58: 1615-1623 (1994).
Amacher, M.C. Nickel, cadmium, and lead. In Methods of
    Soil Analysis: Chemical Methods, Part 3. D. L. Sparks
    (Ed.), Soil Science Society of America, Madison, Wl,
    pp. 739-768(1996).
Amonette, J.E.  Methods for determination  of mineral-
    ogy and environmental availability. In So/7 Mineralogy
    with Environmental Applications. J.B. Dixon and D.G.
    Schulze (Eds.), Soil Science Society of America Book
    Series. Soil Science Society of America, Inc., Madison,
    Wl, pp. 153-197(2002).
Baes, C.F and R.E. Mesmer.  The  Hydrolysis of Cations.
    Wiley Interscience, New York, NY, 489 pp. (1976).
Bailey, E.H., J.F.W. Mosselmans, and S.D. Young. Time-
    dependent surface reactivity of Cd sorbed on calcite,
    hydroxylapatite and humic acid. Mineralogical Maga-
    zine, 69: 563-575 (2005).
Bellanca, A., S. Mauser, R. Neri, and B. Palumbo. Mineralogy
    and geochemistry of Terra Rossa soils, western Sicily:
    insights into heavy metal f ractionation and mobility. Sci-
    ence of the Total Environment, 193: 57-67 (1996).
Benjamin, M.M. and J.O. Leckie. Adsorption of metals at
    oxide surfaces:  Effects  of the  concentrations of ad-
    sorbate and competing metals. In Contaminants and
    Sediments, Volume 2, Analysis,  Chemistry, Biology.
    R.A. Baker (Ed.), Ann Arbor Science Publishers, Ann
    Arbor, Ml (1980).
Buanuam, J., K. Tiptanasup, J. Shiowatana, M. Miro,  and
    E.H. Hansenc. Development of a simple extraction cell
    with bi-directional  continuous flow coupled  on-line to
    ICP-MS for assessment of elemental associations in
    solid samples. Journal of Environmental Monitoring, 8:
    1248-1254(2006).
Carroll, S.A., PA. O'Day, and M. Piechowski.   Rock-wa-
    ter interactions controlling zinc, cadmium, and lead
    concentrations in surface waters and sediments, U.S.
    Tri-state Mining District. 2. Geochemical interpretation.
    Environmental Science and  Technology 32: 956-965
    (1998).
Carroll, S., PA. O'Day, B. Esser, and S. Randall. Speciation
    and fate of trace metals in estuarine sediments under
    reduced and oxidized conditions, Seaplane Lagoon,
    Alameda Naval Air Station (USA). Geochemical Trans-
    actions, 3:81-101  (2002).
Chiarello, R.P, N.C. Sturchio,  J.D. Grace, P. Geissbuhler,
    L.B. Sorensen, L.  Cheng,  and S. Xu. Otavite-calcite
    solid-solution formation at the calcite-water interface
    studied in situ by synchrotron X-ray scattering. Geochi-
    mica et Cosmochimica Acta 61: 1467-1474 (1997).
Christensen,  J.B. and  T.H. Christensen.   Complexation
    of Cd, Ni, and Zn  by DOC in  polluted groundwater:
    A comparison of approaches using  resin  exchange,
    aquifer material sorption, and computer  speciation

-------
    models  (WHAM and  MINTEQA2). Environmental
    Science and Technology 33: 3857-3863 (1999).
Coles, C.A., S.R. Rao, and R.N. Yong. Lead and cadmium
    interactions with mackinawite: Retention mechanisms
    and the role of pH.  Environmental Science and Tech-
    nology 34: 996-1000 (2000).
D'Amore, J.J., S.R. AI-Abed,  K.G. Scheckel, and J.A. Ryan.
    Methods for speciation of metals in soils: A review. Jour-
    nal of Environmental Quality 34:1707-1745 (2005).
Cooper, C. and  J.W.  Morse.   Biogeochemical controls
    on  trace metal cycling  in anoxic  marine sediments.
    Environmental Science and Technology  32: 327-330
    (1998).
Daskalakis, K.D. and G.R. Helz. Solubility of CdS (Gree-
    nockite)  in sulfidic waters at  25  C.  Environmental
    Science and Technology 26: 2642-2468 (1992).
Davis, J.A., C.C.  Fuller, and A.D. Cook. A model for trace
    metal sorption processes at the calcite surface: Adsorp-
    tion of Cd2+ and subsequent solid-solution formation.
    Geochimica  et  Cosmochimica  Acta 51: 1477-1490
    (1987).
DiToro, D.M., J.D. Mahony, D.J. Hansen, J.J. Scott, M.B.
    Hicks, S.M. Mayr, and M.S. Redmond. Toxicity of cad-
    mium  in sediments: The role of acid volatile sulfide.
    Environmental Toxicology and Chemistry 9:1487-1502
    (1990).
Dunnivant, P.M., RM. Jardine, D.L.Taylor, and J.F. McCarthy.
    Cotransport of  cadmium and hexachlorobiphenyl by
    dissolved organic carbon through columns containing
    aquifer material. Environmental Science and Technol-
    ogy 26:360-368 (1992).
Ford, R.G., P.M. Bertsch, and K.J. Farley. Changes in tran-
    sition and heavy metal partitioning during hydrous iron
    oxide aging.  Environmental Science and Technology
    31:2028-2033(1997).
Framson, RE. and J.O. Leckie. Limits of coprecipitation of
    cadmium and ferrous sulfides. Environmental Science
    and Technology 12: 465-469 (1978).
Gibert, O., J. Pablo, J.L. Cortina, and C. Ayora. Evaluation
    of municipal compost/limestone/iron mixtures as filling
    material for permeable reactive barriers for in situ acid
    mine drainage treatment. Journal of Chemical Technol-
    ogy and Biotechnology 78: 489-496 (2003).
Hickey, M.G. and J.A. Kittrick.  Chemical partitioning of
    cadmium, copper,  nickel and zinc  in soils and sedi-
    ments containing high levels of heavy metals.  Journal
    of Environmental Quality 13: 372-376 (1984).
Lai, C.-H., C.-Y. Chen, B.-L. Wei, and S.-H Yeh. Cadmium
    adsorption on goethite-coated sand in the presence of
    humic acid.  Water Research 36: 4943-4950 (2002).
Lindsay, W.L.  Chemical Equilibria in  Soils. John Wiley and
    Sons, New York, NY, 449 pp. (1979).
Martin-Garin, A., P. Van Cappellen, and L. Charlet. Aque-
    ous cadmium uptake by  calcite:  A stirred flow-through
    reactor study. Geochimica et  Cosmochimica Acta 67:
    2763-2774 (2003).
Martinez, C.E. and M.B. McBride. Solubility of Cd2+, Cu2+,
    Pb2+, and Zn2+ in aged coprecipitates with amorphous
    iron hydroxides.  Environmental Science and Technol-
    ogy 32: 743-7'48 (1998).
Martmez, C.E., M.B. McBride, M.T. Kandianis, J.M. Duxbury,
    S.-J.Yoon, and W.F. Bleam. Zinc-sulfur and cadmium-
    sulfur association in metalliferous peats: Evidence from
    spectroscopy, distribution coefficients, and phytoavail-
    ability.  Environmental Science and Technology 36:
    3683-3689 (2002).
McBride, M.B. Chemisorption of Cd2+ on calcite surfaces.
    So/7 Science  Society of America Journal 44: 26-28
    (1980).
O'Day, P.A., S.A. Carroll, and G.A. Waychunas.  Rock-
    water interactions controlling zinc, cadmium, and lead
    concentrations in surface waters and sediments, U.S.
    Tri-state Mining District. 1.  Molecular identification using
    X-ray absorption spectroscopy.  Environmental Science
    and Technology 32: 943-955 (1998).
O'Day, P.A., S.A. Carroll, S. Randall, R.E. Martinelli, S.L. An-
    derson, J. Jelinski, and J.R Knezovich. Metal speciation
    and bioavailability in contaminated estuary sediments,
    Alameda  Naval Air Station, California. Environmental
    Science and Technology 34: 3665-3673 (2000).
Ngiam, L.-S.,  and P.-E. Lim. Speciation patterns of heavy
    metals  in  tropical estuarine anoxic and oxidized sedi-
    ments by  different sequential extraction schemes. Sci-
    ence of the  Total Environment 275:  53-61 (2001).
Papadopoulos, P. and D.L. Rowell. The reaction of cadmium
    with calcium carbonate surfaces. Journal of Soil Sci-
    ence 39: 23-36 C\ 988).
Parkman,  R.H.,  J.M. Charnock, N.D. Bryan, F.R. Livens,
    and D.J. Vaughan.  Reactions of copper and cadmium
    ions in aqueous solution with goethite, lepidocrocite,
    mackinawite, and  pyrite.   American Mineralogist 84:
    407-419 (1999).
Peltier, E., A.L. Dahl, and J.-F. Gaillard.  Metal  speciation
    in anoxic  sediments: When sulfides can be construed
    as oxides. Environmental Science and Technology 39:
    311-316(2005).
Pustisek, N.,  R. Milacic, and M. Veber.  Use of the BCR
    three-step sequential  extraction procedure for  the
    study of the partitioning of Cd, Pb  and Zn in various
    soil  samples.  Journal of Soils and Sediments 1: 25-
    29 (2001).
Pye, K.  Forensic examination of rocks, sediments, soils
    and dusts using scanning electron microscopy and X-
    ray chemical microanalysis. In Forensic Geoscience:
    Principles, Techniques and Applications. K. Pye and
    D.J. Croft (Eds.), Geological Society Special Publica-
    tions, Volume 232, pp.103-122 (2004).
Reimann, C. and P. Caritat. Chemical Elements in the En-
    vironment. Springer-Verlag, Berlin, 398 pp. (1998).
Rudd, T, J.A. Campbell, and J.N. Lester. Use of model
    compounds to elucidate metal forms in sewage sludge.
    Environmental Pollution 50: 225-242 (1988).

-------
Saeki,  K., M. Okazaki, and  S. Matsumoto. The chemi-
    cal phase changes in heavy metals with drying and
    oxidation of the lake sediments. Water Research 27:
    1243-1251 (1993).
Santillan-Medrano, J. and J.J. Jurinak.  The chemistry of
    lead and cadmium in soil: solid phase formation.  So/7
    Science Society of America Proceedings 39: 851 -856
    (1975).
Sery, A., A. Manceau, and G.N. Greaves. Chemical state
    of Cd in apatite  phosphate ores  as determined by
    EXAFS spectroscopy. American Mineralogist 81: 864-
    873(1996).
Shannon, R.D. and J.R. White. The selectivity of a sequen-
    tial extraction procedure for the determination of iron
    oxyhydroxides and  iron sulfides in lake sediments.
    Biogeochemistry 14: 193-208 (1991).
Simpson, S.L., S.C. Apte, and G.E. Batley. Effect of short-
    term resuspension events on the oxidation of cadmium,
    lead,  and  zinc sulfide phases in anoxic  estuarine
    sediments.  Environmental Science and Technology
    34: 4533-4537 (2000).
Stipp, S.L., J.R Hochella, G.A. Parks, and J.O. Leckie. Cd2+
    uptake by calcite, solid-state diffusion, and the forma-
    tion of  solid solution: Interface processes  observed
    with near-surface sensitive techniques (XPS,  LEED,
    and AES).  Geochimica  et Cosmochimica Acta 56:
    1941-1954(1992).
Stipp, S.L., G.A. Parks, O.K. Nordstrom, and J.O. Leckie.
    Solubility-product constant and thermodynamic prop-
    erties for synthetic otavite, CdCO3(s), and aqueous
    association  constants for the Cd(ll)-CO2-H2O system.
    Geochimica et Cosmochimica Acta 57: 2699-2713
    (1993).
Tesoriero, A.J. and J.R  Pankow. Solid solution partition-
    ing of Sr2+,  Ba2+, and Cd2+ to calcite. Geochimica et
    Cosmochimica Acta 60: 1053-1063 (1996).
Tessier, A., P.G.C. Campbell, and M. Bison. Sequential ex-
    traction procedure for the speciation of particulate trace
    metals. Analytical Chemistry 51: 844-851 (1979).
Tonkin, J.W., L.S. Balistrieri, and J.W. Murray. Modeling sorp-
    tion of divalent metal cations on hydrous manganese
    oxide using the diffuse double layer model. Applied
    Geochemistry^: 29-53 (2004).
USE PA. Technology Alternatives  for the  Remediation
    of Soils Contaminated with As, Cd, Cr, Hg, and Pb,
    EPA/540/S-97/500,  U.S. Environmental  Protection
    Agency (1997).
USEPA. National Recommended Water Quality Criteria,
    4304T, Office of Water, Washington DC (2006).
Wang,  R and  A. Tessier.  Cadmium  complexation  with
    bisulfide.  Environmental Science and Technology 33:
    4270-4277(1999).
Wang, Y. and E.J. Reardon. A siderite/limestone reactor to
    remove arsenic and cadmium from wastewaters. Ap-
    plied Geochemistry 16: 1241 -1249 (2001).
Wolke, R.L.  Cadmium.  Chemical and Engineering News
    September 8, p. 120 (2003).
Zachara, J.M., C.E. Cowan, and  C.T Resch. Sorption of
    divalent metals on calcite.  Geochimica et Cosmochi-
    mica Acta 55: 1549-1562 (1991).
Zhao, X., W.H. H6II,  and G. Yun.  Elimination of cadmium
    trace contaminations from  drinking water. Water Re-
    search 36: 851-858  (2002).

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                                                  Lead
                     Richard T. Wilkin, Patrick V. Brady, and Douglas B. Kent
Occurrence and Distribution
Lead  is distributed in low concentrations in  sedimentary
rocks  and soils.  The average concentration of lead in
shales, sandstones, and carbonate rocks is 20, 7, and 9 mg
kg-1, respectively (Turekian and Wedepohl, 1961). Kabatas-
Pendias and Pendias (1984) report background soil con-
centrations of 17-26 mg Pb kg-1 in the U.S. Anthropogenic
enrichment of lead in near-surface soils stems largely from
airborne deposition of particles derived from fossil fuel com-
bustion (e.g., gasoline and coal).  Lead is a common metal
contaminant at hazardous waste sites, especially at battery
crushing and  recycling facilities (USEPA,  1991).  Indeed,
lead is the most commonly recycled metal: roughly 50% of
lead production is secondary lead. Approximately 70% of
world-lead goes to lead-acid storage batteries. Natural lead
enrichment occurs around hydrothermal deposits and base
metal  ores, most frequently as the mineral galena (PbS),
but also as the oxidation products of lead sulfide ores such
as anglesite (PbSO4) and cerussite (PbCO3).

Sources of lead contamination to surface waters and ground
waters include: fall-out of atmospheric dust, industrial and
municipal wastewater effluent, mineral fertilizers and pes-
ticides, lead-based paints, and wastes from the mining,
metallurgical, chemical, and petrochemical industries. Lead
is a widely used non-ferrous metal in the petroleum and
storage battery industries.  In the early 20th century, lead
was frequently used for constructing water pipes and for
the solder used to seal pipe joints, and prior to 1978 lead
carbonates and oxides were common pigment components
in exterior and interior paints. From 1923 to the mid-1980's
tetraethyl lead was used in  the U.S. as an antiknock ad-
ditive  in gasoline, and lead derived from fuel combustion
represented a dominant source of lead to the environment.
Due to a better understanding of the health consequences
stemming from lead exposure, as well as the introduction
of  catalytic converters,  many countries have reduced or
eliminated use of lead additives  in gasoline.  Most  of the
lead produced in the U.S. comes from mines in Missouri,
with other major lead-producing mines in Alaska, Colorado,
Idaho and Montana. The average annual  consumption of
lead in the U.S. from 1997 to 2001 was 1,690,000 metric
tons.

Lead was ranked second on the CERCLA Priority List of
Hazardous Substances in 1999 and 2001 (after arsenic -
#1, and before mercury - #3, vinyl chloride - #4, and PCBs
- #5).  The priority list is prepared by the Agency for Toxic
Substances and Disease Registry and EPA and is based
on the frequency of occurrence of particular contaminants
at National Priorities List  (NPL) sites and their potential
threat to human health.  In absolute numbers, lead is by
far the most common inorganic contaminant found at  Su-
perfund sites.  For example,  in 1996 lead contamination
was found at 460  Superfund  sites, compared  to 306 with
chromium  contamination, 235 with arsenic, 226 with zinc,
224 with cadmium, 201 with copper,  and 154 with mercury
(USEPA, 1996).

Plume Characteristics
The fate of lead in the subsurface is controlled principally
by adsorption at the solid-water interface, precipitation, and
complexation with  organic matter. Lead is strongly retained
in soils and in most situations very little lead  is transported
to surface waters or ground water.  Exceptions  to  this
behavior are low  pH systems or environments with  high
concentrations of  dissolved organic carbon. Tetramethyl
lead, a more soluble and volatile form of lead, may form as
a result of microbial alkylation of inorganic lead in anaerobic
environments. Remedial case studies at lead-contaminated
Superfund sites primarily describe soil cleanup technolo-
gies (U.S.  EPA, 1997). Fewer examples are found where
remediation efforts have targeted lead contamination in
ground water (Morrison and Spangler, 1993).

Remedial Technologies
Technology classes potentially applicable  to  the reme-
diation of  lead-contaminated soils  include containment,
solidification/stabilization, and separation/concentration
(e.g., USEPA, 1997). Containment technologies applied at
metal contamination sites include caps and vertical barriers
to minimize the transport of lead and co-contaminants out
of source  zones.  Solidification/stabilization technologies
are treatment processes  that mix  reactive agents  with
contaminated material to reduce solubility or otherwise limit
contact  between the contaminated material and potential
transport fluids. Amendments such as Portland cement
or phosphate-based compounds are candidates for treat-
ment of lead contamination in soils (e.g., USEPA, 1997).
Separation/concentration methods have also been used for
lead treatment, including ex-situ soil washing and in-situ
soil flushing to physically or chemically reduce contaminant
concentrations to meet site-specific cleanup goals. Ground-
water remediation of lead using the permeable reactive
barrier technology has been explored with some success
in bench-top studies (e.g., Shokes and Moller, 1999).
                                                    11

-------
Regulatory Aspects

Because of the highly toxic effects of lead on biological
systems, treatment standards and concentration limits are
stringent for the discharge  of lead-bearing wastewaters
and ground water. EPA has set the maximum contaminant
level for lead at 0.015 mg L1. Generally, the predominant
routes of exposure for lead are ingestion and inhalation of
lead-bearing aerosols. As will be discussed below, lead is
sparingly soluble in water over a wide range of chemical
conditions. For non-potable water sources, ambient water
quality criteria (AWQC) that are protective of  aquatic life
may serve as alternative cleanup goals.  For lead, current
statutes list both acute and chronic criteria for fresh waters
as 0.065 mg  L1 and 0.0025 mg L1, respectively, for a water
hardness of  100 mg L1 (USEPA, 2006; http://www.epa.
gov/waterscience/criteria/nrwqc-2006.pdf).  Adjustments
to these criteria are to be applied for waters with different
hardness. An example of where this criterion may apply
is a site where contaminated ground water discharges to
surface water.  Cleanup goals for lead in soils at Superfund
sites range from 200 to 500 mg kg-1 (USEPA, 1997).

Geochemistry and Attenuation Processes

Aqueous Speciation

Lead is known to form stable aqueous complexes with OH-,
CI-, CO32-, SO42-, and HS-. In pure water, lead is  mainly
present as Pb2+ below a pH of about 7. With increasing
pH, the species PbOH+,  Pb(OH)2, and Pb(OH)3- become
dominant over Pb2+ (Figure 2.1). Lead carbonate complexes
(PbCO3, Pb(CO3)22-),  lead chloride complexes (PbO,
PbCI2),  lead  sulfate  complexes  (PbSO4,  Pb(SO4)22-),
and  lead sulfide complexes (PbHS+, Pb2S2) are typically
considered in aqueous speciation  modeling efforts (e.g.,
Hem and Durum, 1973; Hem,  1976; Marani et al., 1995;
Pierrard et al., 2002; Rozan et al., 2003). In general, com-
plexation of lead with chloride and sulfate is most important
at near-neutral to low pH; carbonate complexation is most
important at near-neutral to moderately alkaline conditions;
and, hydroxyl species dominate lead complexation at high
pH (Figure 2.2). Inorganic speciation of lead in site-specific
  1.20
                   6     8      10     12
  o.oo
water will depend on pH, total lead concentration, and the
relative and absolute  abundances of the major anions:
chloride, sulfate, and carbonate.

Dissolved  organic carbon  (DOC)  may also form stable
complexes with lead and play an important role in governing
lead mobility in ground-water systems; however, few data
are available and comparatively few attempts have been
made to assess the importance of lead interactions with
DOC. In a recent study of landfill leachate-polluted ground
water containing up to  180 mg DOC L-1, more than 90% of
the total lead in solution was present in DOC complexes
(Christensen et al., 1999). This study also showed that the
default database for MINTEQA2, which contains information
for calculating metal complexation by DOC, was adequate
for predicting the extent of lead complexation by DOC.

Reed et al. (1995) took advantage of lead partitioning to or-
ganic compounds in column-scale soil flushing studies. For
soils contaminated with Pb(ll) (500 mg kg-1 from Pb(NO3)2),
PbSO4 (10,000 mg kg-1), PbCO3 (10,000 mg kg-1), and Pb-
naphthalene (400 mg kg-1), they documented Pb recover-
ies  of 100%, 100%, 100%, and 72%, respectively,  using
0.01 M EDTA as the soil-flushing solution. These results
demonstrate the degree to which lead can be mobilized by
organic ligands such as EDTA.
                                         Pb2*
                                         Pb(OH)*
                                           Pb(OH)2
                                         -Pb(OH)3
                                         Pbcr
                                         PbS04
                                         PbC03
                                         Pb(CO3)22
Figure 2.1  Pb(ll) species distribution in pure water at
            25 C.
Figure 2.2  Species distribution of lead in solution with
            100 mg L1 chloride, 100 mg L1 sulfate, and
            100 mg L1 total inorganic carbon, based on
            thermodynamic data in MINTEQA2 (Allison
            etal.,  1990).


Solubility

An Eh-pH diagram for lead is shown in Figure 2.3. Inspec-
tion of this diagram indicates that at the specified conditions
lead is stable in solids across the stability field of liquid
water.  At low pH  and oxidizing conditions,  lead sulfate
is stable. At near-neutral to moderately alkaline pH, lead
carbonates are stable, and at pH > 12.5 lead hydroxide is
stable. In sulfidic environments, lead sulfide (galena) is
stable over a wide pH range. The Pb(l V) phase, plattnerite,
is stable at moderately alkaline to alkaline pH and at highly
oxidizing  redox potentials.
                                                    12

-------
Solution pH plays a dominant role in governing lead solubil-
ity in aqueous solution. In general, the aqueous solubility of
lead is low at near neutral to alkaline pH. Lead is expected
to be mobile in low pH, oxidizing conditions. Hem and Du-
rum (1973) found that at pH>7, the equilibrium solubility of
lead was below 0.05 mg L1 when Pb(OH)2 and PbCO3 were
assumed to be the solubility-controlling phases. Equilibrium
solubility of greater than 1000 mg L1 lead was estimated
at pH 4 in the absence of any sulfate. Lead is usually not
a metal of concern at mining-related sites where acid mine
drainage is produced. This is because the weathering of
metal sulfides, in addition, to generating acidity also pro-
duces high concentrations of sulfate, which results  in the
precipitation of anglesite (Zanker et al., 2002).

For comparison purposes, the pH-dependent solubilities of
lead carbonate, lead sulfate, and lead sulfide are shown in
Figure 2.4. Lead carbonate is highly insoluble at pH>8, but
can be highly soluble below pH 6. Consequently, acidifica-
tion of a soil or sediment containing lead carbonate may
result in lead mobilization.  Lead sulfate solubility is pH-in-
dependent above pH of  about 2 and the concentration of
Pb(ll) in equilibrium with lead sulfate varies inversely with
the concentration of sulfate. Lead sulfide is highly insoluble
even at low pH (Figure 2.4).
   0.5

  -0.5
                                      Plattnerite
      25'
                       6     8
                         PH
                                  10
                                        12
                                              14
Figure 2.3 Eh-pH diagram for lead (total Pb= 1&s mol-
            al, total C= 10-3 molal, total S=ia3 molal;
            all organic lead complexes are suppressed;
            activity coefficients for all species are set
            equal to 1).


Important  lead-bearing  mineral  phases  include:  lead
hydroxide (Pb(OH)2), cerussite (PbCO3),  hydrocerussite
(Pb3(CO3)2(OH)2), anglesite (PbSO4), galena (PbS), lead
oxide (PbO),  and chloropyromorphite  (Pb5(PO4)3CI)  in
phosphate-bearing systems.  In addition,  plumbojarosite
(Pb05Fe3(SO4)2(OH)6) has been identified as an important
secondary precipitate and  lead sink in weathered mine
wastes  (e.g.,  Hochella  et  al.,  1999).   Thermodynamic
data for most of these phases may be found in a variety
of sources (e.g., see Pierrard et al., 2002, and references
therein). In carbonate and sulfate systems the most favored
mineral species appear to be anglesite, cerussite, and hy-
drocerussite (Lindsay, 1979; Marani et al., 1995).

Lead hydroxide and lead oxide, although predicted to be
stable based on thermodynamic reasoning,  seem  to be
kinetically hindered from precipitating at room temperature
(Marani et al.,  1995). In sulfate-reducing systems, galena
precipitation is thermodynamically and kinetically favored
over a wide range of pH and total  sulfide concentrations
(UhlerandHelz, 1984).

Marani et al. (1995) point out that the reliability of solubility
predictions depends on the choice of the relevant solubility
constants used in modeling studies.  Unfortunately, such
constants are wide ranging for lead. Reasonable agreement
between solubility predictions from equilibrium modeling
and filterable lead concentrations measured in  aged soil-
water systems was obtained only with a critical selection
of solid phases in the modeling and  by appreciating kinetic
aspects of the Pb-H2O system  (Marani et al., 1995).
                                                         1.0E+06
                                                         1.0E+00
.
>  1.0E-06
13
is

= 1.0E-12
D.
                                                          1.0E-18
                                                          1.0E-24
                                                                      PbCO3 (ETIC=1 mm)
                                                                    PbS04(SS04=100mm)
                                                                               PbS (2H2S=1 mm)
                                                                                 6       8

                                                                                     PH
                                                                                                 10
                                                                                                          12
Figure 2.4 Pb(ll) activity in equilibrium with PbCO3 (at
            total inorganic carbon equal to 0.001 molal),
            PbSO4 (at total sulfate equal to 0.1 molal),
            and PbS (at total sulfide equal to 0.001
            molal). The solubility trend can be com-
            pared to the MCL for lead of 0.015 mg L1
            or an activity of~7.2x10~8 m assuming ideal
            behavior.
Lead  phosphate minerals appear to be highly  insoluble
lead-bearing phases and  remediation strategies for sta-
bilizing lead-contaminated soils have taken advantage of
this behavior (e.g., Ruby et al., 1994; Zhang et al., 1997).
Hydroxyapatite and sodium phosphate monobasic have
been  used as a source of soluble phosphate to amend
                                                     13

-------
lead-contaminated soils.   Reaction  between labile lead
phases and dissolved phosphate is rapid over a wide range
of pH and P/Pb molar ratios and results in the formation of
insoluble chloropyromorphite (Zhang and Ryan, 1999).

Adsorption

Adsorption of trace metals, such as lead onto oxide sur-
faces, has  been well characterized  in lab-based studies
(e.g., Hayes and Leckie,  1986).  Adsorption at  mineral
surfaces results from a set of chemical reactions between
lead and various surface sites (Dzombak and Morel, 1990).
Most of these reactions involve the  release of H+  ions,
which accounts for the  strong dependence of adsorption
on pH. Hydrous ferric oxide (HFO) is of particular interest
because it is found at many contaminated sites and could
play a major role in governing  the mobility of lead, other
metals, and metalloids (e.g., Trivedi et al., 2003; Dyer et al.,
2003). Lead adsorbs more strongly onto HFO compared to
most other divalent metal ions (Dzombak and Morel, 1990);
the same is true  for  other ferric oxides, hydrous oxides,
aluminum oxides, oxyhydroxides, clay minerals, and poorly
ordered Fe- and Al-containing hydroxypolymer coatings on
natural aquifer sediments  (Sposito,  1984; Coston et  al.,
1995; O'Reilly and Hochella, 2003).

Long-term  lab studies  typically indicate that adsorption
occurs in two steps: i) rapid initial surface sorption or  ex-
change  followed by, ii)  continued but slow metal uptake.
Ainsworth et al. (1994) reported on long-term aging studies
of lead onto HFO. Consistent with previous investigations,
they found that the adsorption of lead onto HFO increased
sharply from 0 to 100% as pH increased from 3 to 6. The
pH  at which 50% of the  lead  was associated  with  the
HFO surface was 4.7. The sorption behavior of lead onto
HFO was found to be independent of time from 0 to 21
weeks of aging. In addition, desorption of lead from aged
HFO exhibited no hysteresis with the adsorption pH edge
developed from non-aged  HFO. Ainsworth et al. (1994)
concluded that the lead adsorption-desorption process is
completely reversible with time and that there was no indi-
cation that lead was being incorporated into the HFO solid
during aging. The process of recrystallization or aging may
in fact result in the net loss of available sorption sites due
to surface area reductions and  may drive lead desorption
(e.g., Ford et al., 1997).

Results  of a series of laboratory experiments,  which in-
cluded studies of  lead transport through columns packed
with soil, were consistent with these concepts. Experiments
showed that lead mobility  decreased with increasing pH;
lead adsorption  resulted in decreases in  pH,  which, in
turn, increased lead mobility; and, lead adsorption onto the
soil was fast and reversible on the time-scale of transport
(Papini etal., 1999).

Molecular studies of lead sorption onto hydrous ferric  ox-
ide show that Pb(ll) ions associate with the iron hydroxide
surface  mainly as inner-sphere complexes (Trivedi et  al.,
2003). For most of the iron oxides, edge-sharing bidentate
complexes are dominant at pH>5 over a wide range of  ad-
sorbate concentrations (e.g., Bargar et al., 1997; Manceau
etal., 1992). Rouff etal. (2004) reached similar conclusions
regarding lead sorption at the calcite-water interface. How-
ever, at higher initial concentrations of lead (4-12 mg L~1),
precipitation of lead carbonate dominates lead partitioning
in the solid phase (Rouff et al., 2004).

Although lead adsorption in laboratory-based studies  may
be completely reversible, uptake of lead in natural systems
is often substantially irreversible.  Coughtrey et al. (1986)
reviewed soil measurements and suggested that only 50%
of lead in soil was exchangeable. Others have  noted sub-
stantially lower exchangeable fractions (see, e.g., Wang et
al.,  1995; Brady etal., 1999).

Adsorption of iron can influence  the mobility of Pb even
in the presence of strong complexing ligands  like EDTA.
Results of a transport experiment conducted  in a mildly
acidic, quartz-sand aquifer showed that Pb was displaced
from EDTA complexes by Fe(lll) dissolved from aquifer sedi-
ments over short transport distances (Davis et al., 2000).
Even though Pb forms strong complexes with EDTA, strong
adsorption of Pb at oxide surfaces enhances the thermody-
namic driving force for displacement from EDTA complexes
at mildly acidic pH values. However, decreasing solubility
of Fe oxides with increasing  pH decreases the affinity of
the displacement reaction with increasing pH so that at pH
values greaterthan 8 the reaction is unfavorable (Xue et al.,
1995; Nowack et al., 2001). Lead was also displaced from
EDTA complexes by Zn desorbed from Zn-contaminated
sediments (Davis et al., 2000). These types of reactions
may limit the extent to which  complexing ligands enhance
the transport of Pb in contaminated systems.

In reducing systems, adsorption of lead to iron sulfide sur-
faces is possible.  Jean and  Bancroft (1986) investigated
the pH-dependent adsorption behavior of Pb  on several
iron sulfide minerals and found that the pH at which 50%
of the lead was associated with the pyrite (FeS2) surface
was about 6.0. They suggested that the observed adsorp-
tion behavior is controlled by the hydrolysis of Pb2+ ions,
whereby hydrolyzed species sorb directly on sulfide surface
groups as a monolayer.

Lead adsorption onto or co-precipitation with amorphous
FeS may be extensive in reducing systems. Experimental
studies suggest that a significant fraction of the Fe in freshly
precipitated FeS may be replaced rapidly by metal ions like
Pb that form less soluble sulfides (Phillips and Kraus, 1965;
Caletka et al., 1975; Coles et al.,  2000).

Redox Chemistry

In natural systems lead is present in the +2 oxidation state
over relevant conditions of pH and oxidation-reduction po-
tential. Reduction of Pb2+ to metallic  lead is expected to
occur at redox potentials below the stability field of water at
pH<6.  The oxidized form of lead  (Pb(IV)) is  not expected
in air-saturated solutions based  on thermodynamic  rea-
soning. However, the mineral plattnerite (PbO2) occurs in
some natural systems and is associated with other oxida-
tion products such as cerussite  and  pyromorphite (e.g.,
Yeates and Ayres, 1892; see Figure 2.3).  In general, the
geochemical transport processes of lead are not directly
                                                    14

-------
tied to redox conditions. However, because lead may form
stable precipitates with redox-sensitive elements such as
sulfur, lead mobility is indirectly tied to redox conditions.  In
sulfate-reducing systems, lead is expected to form insoluble
PbS precipitates.  In moderately reducing but non-sulfidic
systems, however,  reductive dissolution of hydrous ferric
oxides that  contain adsorbed lead  could result  in lead
mobilization.

Colloidal Transport

The transport of lead in particulate forms can be significant
in ground water and surface water systems. Colloids are
generally considered to be particles with diameters less
than 10 micrometers (Stumm and Morgan, 1996). Colloidal
particles can be present as mineral or organic forms. For
the special case of lead sorbed to colloidal hydrous ferric
oxides, changes in geochemical  regimes may either favor
increased lead sorption or desorption. Increases in lead
sorption may result from increases in pH or Eh. Conversely,
decreases in pH or Eh may result in lead remobilization.
Decreases in the ionic strength of the aqueous phase can
enhance colloidal stability  and  promote  lead  transport,
whereas increases in ionic strength  can promote colloid
aggregation  and removal from the aqueous phase.

Site Characterization

Lead mobility in ground water is governed by the total con-
centration of lead, the distribution of lead species in water,
and the nature of lead partitioning in the solid phase. The
development of site-conceptual models for predicting the
long-term fate of lead at a contaminated site will  require
information on the distribution and concentration of lead in
the aqueous phase and the solid phase. Table 2.1 indicates
possible  natural attenuation and  mobilization  pathways
for lead.  Details of the types of analytical measurements
that  may  be conducted on sampled ground water  and
aquifer sediments to assist in identifying  the attenuation
mechanism(s) are discussed in the following paragraphs.
                       Aqueous Measurements

                       Quantitative measurement of lead concentrations in aque-
                       ous solutions is typically carried  out using inductively
                       coupled plasma atomic emission spectroscopy (ICP-AES),
                       inductively coupled plasma mass spectroscopy (ICP-MS),
                       or atomic absorption spectroscopy (AAS).  Input data to
                       geochemical codes (e.g., MINTEQA2, PHREEQC, EQ3/6)
                       for determining  aqueous speciation  also  require, at a
                       minimum, concentrations of major anions, major cations,
                       dissolved organic carbon, temperature, and pH. In addi-
                       tion, while regulatory requirements stipulate that unfiltered
                       ground-water samples be analyzed to support regulatory
                       decisions at a contaminated site, it may be necessary to
                       also collect filtered samples to help define the process(es)
                       controlling contaminant mobility. The use of 0.45 urn pore
                       size filters is common as an arbitrary cutoff point to  dif-
                       ferentiate between dissolved and  particulate phases in
                       water samples.  However, caution is recommended when
                       using this approach, particularly for Fe and Al and other
                       elements that may be associated with Fe or Al particles
                       (including Pb) that could pass through 0.45 urn filters. The
                       use of filters with pore sizes less than 0.1 um will generally
                       provide a better assessment of the dissolved vs. particulate
                       load in ground water.

                       Solid Phase Measurements

                       The implementation of an analytical approach to identify
                       lead speciation in aquifer sediments is a challenging pro-
                       cess. The accuracy of the analytical finding is dependent
                       on the method of sample collection/preservation and  the
                       tools used to identify the mechanism of lead partitioning. It
                       is recommended that the analytical protocol be designed to
                       address the potential redox sensitivity of the solid phase(s)
                       to which  lead may be partitioned (e.g.,  sulfides in reduced
                       sediments).   Tools to evaluate the mechanism of lead
                       solid phase partitioning range in complexity from relatively
                       simple chemical  extractions to advanced spectroscopic
                       techniques.
Table 2.1  Natural attenuation and mobilization pathways for lead.
    Attenuation Processes
    Mobilization Processes
         Characterization Approach
 Precipitation of insoluble car-
 bonates, sulfides, sulfates, and
 phosphates. In general, pH>8
 will drive precipitation reactions
 resulting in Pb concentrations
 to below the MCL.
Dissolution of carbonates at
low pH; oxidative dissolution of
sulfides at low pH and high Eh;
complexation/stabilization in the
presence of DOC.
Evaluation of lead speciation in the aqueous
phase. Determination of total Pb in the solid
matrix. Evaluation of solid phase partition-
ing using sequential extraction methodologies.
Evaluation of long-term capacity.
 Sorption to iron hydroxides
 (reversible), organic matter,
 carbonates, sulfides (pH>5).
Desorption at low pH; com-
plexation/stabilization in the
presence of DOC.  Reductive
dissolution of iron hydroxides.
Evaluation of lead speciation in the aqueous
phase. Determination of total Pb in the solid
matrix. Evaluation of solid phase partition-
ing using sequential extraction methodologies.
Batch and column testing to determine Pb up-
take capacity of site-specific aquifer materials
with variable geochemical conditions.
                                                    15

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The total concentration  of lead in soils, sediments, and
aquifer materials may be determined by X-ray fluorescence
spectroscopy (XRF),  or by  ICP-AES after  digestion in
mineral acids. A variety of digestion or extraction methods
can be found in the literature (Amacher, 1996). Neutron
activation analysis is not commonly employed due to the
scarcity of neutron sources required to irradiate the sample.
X-ray fluorescence is the most attractive  approach due
to the relative ease of sample preparation, which may be
conducted with  the sample in its original state or following
fusion with  lithium  metaborate. When combined with the
determination of other major  or trace elements in the solid
sample, this provides an initial step for assessing possible
association of lead with  various solid phase components.
This type of analysis can be conducted on the bulk sample
as well as at a microscopic level using wavelength (electron
microprobe) or energy dispersive spectroscopy coupled to a
scanning or transmission electron microscope. Microscopic
examination allows one to better differentiate whether lead
may be distributed  across a number of  different mineral
phases within the solid sample or primarily associated with
a discrete phase.  There are limitations to this approach
(Pye, 2004), a significant one being that the analysis does
not  necessarily provide  unique mineral identification ne-
cessitating the collection of supporting mineralogical and
chemical data.

More  detailed  information  on the specific partitioning
mechanism(s)  controlling lead solid phase speciation is
typically required to adequately support  site assessment
for potential  reliance on natural attenuation as part of a site
remedy. There have been many applications of sequential
extraction schemes to assess the speciation of solid phase
lead (e.g., Tessier et al., 1979; Harrington et al.,  1998;
Sutherland,  2002;  Zanker et al.,  2002; Buanuam et al.,
2006).  As discussed in the cited reports, sequential extrac-
tion methods provide a useful tool  to assist in determining
the chemical speciation of trace metals in soils/sediments,
but  essentially  all documented methods show analytical
limitations in selectively  extracting lead and other metals
associated with specific solid components (e.g., Scheckel et
al., 2003). Design and application of extraction procedures
should take into account the chemical behavior of lead
relative to potential analytical bias that may be introduced
by the extraction chemistry.  For example, due to its low
solubility, lead carbonate would be anticipated to form dur-
ing extractions conducted using solutions with pH buffered
by an excess of dissolved bicarbonate-carbonate.

Where feasible, it is  recommended  that complimentary
analytical techniques  be employed to confirm the  accu-
racy of lead speciation (e.g.,  O'Day et al., 1998; O'Day et
al., 2000; Carroll et al., 2002; D'Amore et al., 2005) or the
accuracy of the extraction of a targeted phase(es) for a
given extractant (e.g., Shannon and White, 1991; Ngiam
and Lim, 2001; Peltier et  al., 2005). As an example, Peltier
et al. (2005) have demonstrated that a common extraction
method employed to target metals associated with easily
reducible iron (hydr)oxides may also dissolve iron sulfides
that may be present.  The results from this analysis may
lead to misidentification  of a lead association  with iron
(hydr)oxides, resulting in the development of a conceptual
site model that misrepresents the site-specific attenua-
tion process. Under reducing conditions, it is also critical
that aquifer sediments be sampled and processed in a
manner that prevents exposure to oxygen prior to extrac-
tion in order to limit oxidation  of reduced minerals (e.g.,
iron sulfides) that may host lead. Lead associated with a
sulfidic phase in sediments has been shown to repartition
to more extractable phases upon oxidation  (Saeki et al.,
1993; Cauwenberg and Maes, 1997). Determination of the
host mineral phase(es) dissolved for each extraction step
is recommended, along with the use of surrogate Pb-bear-
ing phases spiked into the sediment to confirm  accuracy
of the procedure (e.g.,  Rudd et al., 1988). The  choice of
appropriate lead surrogate phases would be governed by
site-specific geochemical conditions or characterization of
the mineralogy of the aquifer sediment.

Pb Isotopes

Information about the source of  lead contamination at a
given  site can be gained using isotopic analysis, particu-
larly by examining the  ratios 206Pb/204Pb, 207Pb/204Pb, and
208Pb/204Pb (Emmanuel and Erel, 2002). As an  example,
Chow and Johnstone (1965) reported isotope ratios of lead
extracted from gasoline (purchased in 1965).  Their data
demonstrated the similarity of the isotope compositions of
lead in gasoline,  airborne particles in Los Angeles, and
snow  from Lassen Volcanic Park.  The study showed that
lead in the air and snow in California in the 1960's originated
from fuel combustion exhaust. Gulson et al. (1981) analyzed
the lead isotopic composition of soils in South Australia to
identify the source  of lead contamination.  By analyzing
and comparing isotope ratios they determined that orchard
sprays, power stations, and smelters were not the principal
source of lead  contamination,  rather lead contamination
was again derived from tetraethyl lead in gasoline.

Long-term Stability  and Capacity

The stability of attenuated lead will depend on the temporal
stability of site geochemical conditions. For example, if lead
attenuation follows a lead sulfide precipitation pathway, then
long-term stability of attenuated lead may depend on the
persistence of reducing conditions.  It is therefore critical to
understand attenuation mechanism(s) so that geochemical
triggers for remobilization can be anticipated and incorpo-
rated into evaluations of long-term monitoring data. For any
proposed and identified attenuation mechanism, there will
exist possible scenarios whereby remobilization could occur
(i.e., changes in pH or Eh).  It will be essential to explore the
likelihood of such  changes in prevailing site geochemistry
and the sensitivity of the attenuation pathway to changes
in the prevailing geochemical conditions.

Quantifying the attenuation capacity (as defined in Volume
1) will also require an understanding of the specific attenu-
ation pathway(s).  Attenuation capacity, for example, could
be related to the extent that pH is buffered, the availability
of sorptive sites in  aquifer materials, or to the supply of
electron donors needed to sustain microbially mediated re-
dox conditions.  For any proposed attenuation mechanism,
                                                     16

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there will be assumptions built into capacity estimations,
so it is recommended that uncertainty analysis accompany
capacity estimates.

Tiered Analysis

Determination of the viability of lead remediation in ground
water via monitored natural attenuation will depend upon
proper assessment of contaminant loading to the aquifer
and prevailing geochemistry  and mineralogy within the
contaminant plume and the down gradient zone prior to
the point(s) of compliance.  MNA may not be appropri-
ate as a site remedy for lead contamination in  acidic to
circum-neutral pH, highly oxidizing, and/or DOC-rich envi-
ronments. The goal of site assessment is to demonstrate
the process(es) controlling lead sequestration onto aquifer
solids and the long-term stability of solid phase lead as a
function of existing and anticipated ground-water chemistry.
The following tiered analysis structure for site characteriza-
tion provides an approach to evaluate candidate sites and
define the potential limitations of MNA as part of a remedy
for ground-water cleanup.

Tier I - Site characterization under Tier I will involve dem-
onstration that the plume is static  or shrinking, has  not
reached  compliance boundaries,  and  does  not impact
existing water supplies.  Once this is  established through
ground-water characterization,  evidence is collected  to
demonstrate  Pb partitioning to aquifer solids within the
plume. Rapid movement of contaminants along preferred
flow paths in the unsaturated and saturated zones can  by
induced by hydrologic events such as heavy rains. It will  be
important to determine that such hydrogeologic features  do
not result in contaminants bypassing zones where natural
attenuation is occurring.  If natural attenuation processes
are active throughout the plume, then there should be  an
observed increase in solid phase  concentrations within
regions of the plume with higher aqueous concentrations,
e.g., near the source term.  This field partitioning data
may be supplemented by geochemical modeling that  in-
corporates measured water chemistry (e.g., pH, Eh, and
major ion chemistry) throughout the plume to assess the
potential  for solubility control by a lead precipitate such
as  a  carbonate/phosphate or sulfide phase.  Identifica-
tion of active sequestration to prevent  lead migration in
ground-water provides justification for proceeding to Tier
II characterization efforts.

Tier II- Under Tier II, the apparent rate and mechanism(s)
of attenuation are determined. Estimates of a site attenu-
ation  rate(s) can be  assessed via  a  well transect along
the ground-water flow path.  In addition, time-series data
may be collected at one or more monitoring points within
the plume. This  information will allow assessment of the
relative timescales for contaminant immobilization and fluid
transport and determination of whether remediation objec-
tives can be met  within the required regulatory time frame.
In addition, the mechanism(s) for attenuation  need to  be
identified under this  stage of site characterization. This
effort may require determination of the chemical speciation
of aqueous and solid  phase Pb, which may be approached
according to the  following scheme:
 1)   Determination of lead solution  speciation via di-
     rect analytical measurements in combination with
     speciation calculations based  on characterized
     ground-water chemistry;
 2)   Calculation of the saturation state of ground water
     relative to measured  aqueous chemistry compli-
     mented by the possible  isolation of discrete Pb
     mineral phases via density separations (or other
     schemes)  in regions  of the aquifer with highest
     solid phase concentrations;
 3)   Determination of aquifer mineralogy to determine
     the relative abundance of components with docu-
     mented capacity for Pb sorption (e.g., Amonette,
     2002);
 4)   Identification of lead association(s)  with the
     various solid phase components of aquifer solids
     through combination of chemical extractions with
     microscopic/spectroscopic confirmation  of phase
     associations, and;
 5)   Demonstration of concurrence  between the site
     conceptual model and mathematical model(s) that
     describe lead removal mechanism(s).
It is recommended that identification of lead chemical spe-
ciation in aqueous and solid matrices be conducted  using
samples collected in a manner that  preserves the in-situ
distribution of dissolved lead and mineralogy and prevents
loss of lead from aqueous  samples (e.g., due to oxidation
and  precipitation of ferrous iron in anoxic  ground water).
The demonstration of concurrence between conceptual and
mathematical models describing lead transport will  entail
development of site-specific parameterization of the chemi-
cal processes controlling lead solid phase partitioning.

Tier III - Once  the partitioning mechanism(s) have been
identified for the site, the subsequent characterization effort
under Tier III will involve determination of  the stability of
immobilized Pb and the  capacity of the aquifer to sustain
continued uptake.  It is recommended that the stability of
immobilized Pb be tested based on the anticipated evolution
of ground-water chemistry concurrent  with plume shrinkage.
For example, changes in ground-water pH can exert  a sig-
nificant influence on Pb adsorption or precipitate solubility.
Therefore, it is recommended that sediment leach tests be
conducted to characterize the magnitude of Pb mobilization
as a function of pH for a ground-water chemistry representa-
tive of site conditions. It is recommended that the capacity
for Pb uptake onto aquifer solids be determined relative to
the specific mechanism(s) identified in Tier II. For example,
if site characterization under Tier II indicated that precipita-
tion of Pb sulfide due to microbial degradation of organic
compounds coupled with sulfate reduction occurs within the
aquifer, then it is recommended that  the mass distribution
of organic carbon and sulfate to support this reaction within
the  aquifer be determined. This site-specific capacity can
then be compared to  Pb mass  loading within the plume
in order to assess the longevity of the natural attenuation
process. If site-specific  tests demonstrate the stability of
immobilized Pb and sufficient capacity within the aquifer to
sustain Pb attenuation, then the site characterization effort
                                                     17

-------
can progress to Tier IV. For cases where contaminant stabil-
ity is sufficient but aquifer capacity is insufficient for capture
of the entire plume, then a determination of the benefits of
contaminant source reduction may be necessary.

Tier IV - Finally, under Tier IV a monitoring plan is estab-
lished along with contingency plans in the event of MNA
failure.  It is recommended that the monitoring plan be
designed to establish both continued plume stability and to
identify changes in ground-water chemistry that  may lead
to re-mobilization of attenuated  Pb. The specific chemical
parameters to be monitored will include those identified un-
der Tier III that may halt Pb partitioning to aquifer sediments
and/or result in solubilization of  either discrete Pb precipi-
tates or aquifer minerals that sequester Pb from ground
water. For example, solution phase parameters that could
alter either Pb precipitation or adsorption include increases
in soluble organic carbon  in combination with changes in
ground-water pH. In contrast, the concentration of dissolved
iron or sulfate may indicate the dissolution of an important
sorptive phase within the aquifer (e.g., reductive dissolution
of iron oxides or oxidative dissolution of su If ides).  Changes
in these parameters may occur prior to observed changes
in solution Pb and, thus, serve  as monitoring triggers for
potential MNA failure. In this instance, a  contingency plan
can be implemented that incorporates strategies to arrest
possible plume expansion beyond compliance boundaries.
Possible  strategies to prevent  plume expansion include
pump and treat operations, installation of reactive barriers
to enhance uptake capacity perpendicular to the direction
of plume advance, or enhancement of natural attenuation
processes within the aquifer through the injection of soluble
reactive components.

References

Ainsworth, C.C., J.L. Pilon,  PL. Gassman,  and W.G. Van der
    Sluys. Cobalt, cadmium, and lead sorption to hydrous
    iron oxide: residence time effect. So/7 Science Society
    of America 58: 1615-1623 (1994).
Allison, J.D., D.S. Brown,  and K.J. Novo-Gradac.  MINT-
    EQA2/PRODEFA2, A Geochemical Assessment Model
    for Environmental Systems, Version 3.0, U.S. Environ-
    mental  Protection Agency, Athens, GA (1990).
Amacher, M.C. Nickel, cadmium, and lead. In Methods of
    Soil Analysis: Chemical Methods,  Part 3. D.L. Sparks
    (Ed.), Soil Science  Society  of America, Madison, Wl,
    pp. 739-768(1996).
Amonette, J.E.  Methods for determination of mineralogy
    and environmental availability. In So/7 Mineralogy with
    Environmental Applications. Dixon  J. B. and  D.  G.
    Schulze (Eds.), Soil Science Society of America Book
    Series.  Soil Science Society of America, Inc., Madison,
    Wl, pp. 153-197(2002).
Bargar, J.R., G.E. Brown,  and G.A. Parks. Surface com-
    plexation of Pb(ll) at oxide-water interfaces: II. XAFS
    and bond-valence determination of mononuclear Pb(ll)
    sorption products and surface functional groups on iron
    oxide surfaces. Geochimica et Cosmochimica Acta 61:
    2639-2652(1997).
Brady P.V., B.P Spalding,  K.M. Krupka,  R.D. Waters, P.
    Zhang, D.J. Borns, and W.D. Brady. Site screening and
    technical guidance for monitored natural attenuation at
    DOE sites. Sandia National Laboratories, Albuquerque,
    New Mexico. SAND99-0464 (1999).
Buanuam, J.,  K. Tiptanasup, J. Shiowatana, M. Miro, and
    E.H. Hansenc. Development of a simple extraction cell
    with bi-directional continuous flow coupled on-line to
    ICP-MS for assessment of elemental associations in
    solid samples. Journal of Environmental Monitoring 8:
    1248-1254(2006).
Caletka, P.., M.Tymple, and P. Kotas. Sorption properties of
    iron(ll) sulphide prepared by the sol-gel method. Journal
    of Chromatography 111: 93-104 (1975).
Cauwenberg,  P. and  A. Maes. Influence  of oxidation on
    sequential chemical extraction of dredged river sludge.
    International Journal of Environmental Analytical Chem-
    istry 68: 47-57 (1997).
Chow, T.J. and M.S. Johnstone. Lead isotopes in gasoline
    and aerosols  of  Los Angeles basin.  Science 147:
    502-503(1965).
Christensen, J.B., J.J. Botma, and T.H. Christensen.  Com-
    plexation of Cu and Pb by DOC in polluted groundwater:
    A comparison of experimental data and predictions by
    computer speciation models (WHAM and MINTEQA2).
    Water Research 33: 3231-3238 (1999).
Coles, C.A., S.R. Rao, and R.N. Yong. Lead and cadmium
    interactions with mackinawite: Retention mechanisms
    and role of pH. Environmental Science and Technology
    34:996-1000(2000).
Coston, J.A.,  C.C. Fuller, and J.A. Davis. Pb2+ and Zn2+
    adsorption by a natural aluminum- and  iron-bearing
    surface coating on  an aquifer sand. Geochimica et
    Cosmochimica Acta 59: 3535-3547 (1995).
Coughtrey P.J., D. Jackson, C.H. Jones, P. Kane, and M.C.
    Thorne.  Radionuclide  distribution  and transport  in
    terrestrial  and aquatic systems. Vol. 6., A.A. Balkema
    (1986).
D'Amore, J.J., S.R. AI-Abed, K.G. Scheckel, and J.A. Ryan.
    Methods for speciation of metals in soils: A review. Jour-
    nal of Environmental Quality 34:1707-1745 (2005).
Davis, J.A., D.B. Kent, J.A. Coston, K.M. Hess, and J.L.
    Joye.  Multispecies reactive transport experiment in
    an  aquifer with spatial variable chemical conditions:
    Experimental design and observed tracer movement.
    Water Resources Research 36: 119-134 (2000).
Dyer, J.A., P. Trivedi, N.C. Scrivner, and D.L. Sparks. Lead
    sorption  onto ferrihydrite.  2. Surface complexation
    modeling. Environmental Science and Technology 37:
    915-922 (2003).
Dzombak, D.A. and F.M.M. Morel.  Surface Complexation
    Modeling: Hydrous Ferric Oxide. John Wiley and Sons,
    New York, NY  (1990).
Emmanuel, S. and Y Erel. Implications from concentrations
    and isotopic data for Pb partitioning processes in soils.
                                                    18

-------
    Geochimica  et Cosmochimica  Acta 66:  2517-2527
    (2002).
Ford, R.G., P.M. Bertsch, and K.J. Farley. Changes in tran-
    sition and heavy metal partitioning during hydrous iron
    oxide aging.  Environmental Science and  Technology
    31:2028-2033(1997).
Gulson, B.L., K.G.Tiller, K.J. Mizon, and R.H. Merry, 1981.
    Use of lead isotopes in soils to  identify the source of
    lead contamination near Adelaide, South Australia.
    Environmental Science and Technology 15: 691 -696.
Harrington, J.M.,  M.J. Laforce, W.C. Rember, S.E. Fendorf,
    and R.F Rosenweig.  Phase associations and  mobi-
    lization of iron and trace elements in  Coeur d'Alene
    Lake,  Idaho.  Environmental Science and  Technology
    32:650-656(1998).
Hayes, K.F and J.O. Leckie. Mechanism of lead ion adsorp-
    tion at the goethite-water interface. In Geochemical
    Processes at Mineral Surfaces. J.A. Davis and K.F.
    Hayes (Eds.), American Chemical Society, Washington
    DC, 114-141  (1986).
Hem, J.D.  and W.H. Durum. Solubility and occurrence of
    lead in surface water.  Journal of the American Water
    Works Association 65: 562-568  (1973).
Hem, J.D.  Geochemical controls on lead  concentrations
    in stream water and sediments.  Geochimica et Cos-
    mochimica Acta 40: 599-609 (1976).
Hochella,  M.F, J.N. Moore, U. Golla, and A. Putnis.  A
    TEM study of samples from acid mine drainage sys-
    tems:  Metal-mineral association with implications for
    transport.  Geochimica et Cosmochimica Acta  63:
    3395-3406(1999).
Jean, G.E. and G.M. Bancroft.  Heavy metal adsorption by
    sulphide mineral  surfaces.  Geochimica et Comochi-
    mica Acta 50: 1455-1463 (1986).
Kabata-Pendias,  A.  Trace Elements in Soils  and Plants.
    CRC Press, Boston, MA (1984).
Lindsay, W.L.  Chemical Equilibria in  Soils. John Wiley and
    Sons,  New York, NY (1979).
Manceau,  A., L.  Charlet, M.C. Boisset,  B. Didier, and L.
    Spadini. Sorption and speciation of heavy metals on
    hydrous iron and manganese oxides: From microscopic
    to  macroscopic.  Applied Clay Science  7: 201-203
    (1992).
Marani, D., G. Macchi, and M. Pagano.  Lead precipitation
    in the presence of sulphate and carbonate: Testing
    of thermodynamic predictions.  Water Research  29:
    1085-1092(1995).
Morrison,  S.J. and  R.R. Spangler. Chemical  barriers for
    controlling groundwater contamination.  Environmental
    Progress 12: 175-181  (1993).
Ngiam, L.-S. and P.-E. Lim. Speciation  patterns of heavy
    metals in tropical estuarine anoxic and oxidized sedi-
    ments by different sequential extraction schemes. Sci-
    ence of the Total Environment 275:  53-61  (2001).
Nowack, B.,  FG. Kari, and F Krueger. The remobilization
    of metals from iron oxides and sediments by metal-
    EDTA complexes.  Water, Air, and Soil Pollution 125:
    243-257(2001).
O'Day, PA.,  S.A. Carroll, and G.A. Waychunas.  Rock-
    water interactions controlling zinc,  cadmium, and lead
    concentrations in surface waters and sediments, U.S.
    Tri-state Mining District. 1.  Molecular identification using
    X-ray absorption spectroscopy. Environmental Science
    and Technology 32: 943-955 (1998).
O'Day, PA., S.A. Carroll, S. Randall, R.E. Martinelli, S.L. An-
    derson, J. Jelinski, and J.P Knezovich. Metal speciation
    and bioavailability in contaminated estuary sediments,
    Alameda Naval Air Station, California. Environmental
    Science  and Technology 34: 3665-3673 (2000).
O'Reilly, S.E. and M.F. Hochella. Lead sorption efficiencies
    of natural and synthetic Mn and Fe-oxides. Geochimica
    et Cosmochimica Acta 67: 4471-4487 (2003).
Papini, M.P, YD. Kahie,  B. Troia,  and M. Majone.  Ad-
    sorption  of lead at variable pH onto  a natural porous
    meduim: Modeling of batch and column experiments.
    Environmental Science and Technology 33:4457-4464
    (1999).
Peltier, E., A.L.  Dahl, and J.-F Gaillard.  Metal  speciation
    in anoxic sediments: When sulfides can  be construed
    as oxides. Environmental Science  and Technology 39:
    311-316(2005).
Phillips, H.O. and  K.A. Kraus.  Adsorption  on inorganic
    materials VI. Reaction of  insoluble sulfides with metal
    ions in aqueous media.  Journal of  Chromatography
    17: 549-557 (1965).
Pierrard, J., J. Rimbault, and M. Aplincourt.  Experimental
    study and modeling of lead solubility as a function of
    pH in mixtures of ground waters and cement waters.
    Water Research 36: 879-890 (2002).
Pye, K.  Forensic examination of rocks, sediments, soils
    and dusts using scanning electron microscopy and X-
    ray chemical microanalysis. In  Forensic Geoscience:
    Principles, Techniques and Applications.  K. Pye and
    D.J. Croft (Eds.), Geological Society Special Publica-
    tions, Volume 232, pp.103-122 (2004).
Reed, B.E., R.E. Moore, and S.R. Cline.  Soil flushing of a
    sandy loam contaminated with Pb(ll), PbSO4 (s), PbCO?
    (s), or Pb-naphthalene: Column results. Journal of Soil
    Contamination 4: 243-267 (1995).
Rouff, A.A., E.J. Elzinga, R.J.  Reeder, and N.S. Fisher.  X-
    ray absorption spectroscopic evidence for the formation
    of Pb(ll)  inner-sphere  adsorption complexes and pre-
    cipitates at the calcite-water interface. Environmental
    Science  and Technology 38: 1700-1707 (2004).
Rozan, T.F, G.W. Luther, D. Ridge, and S. Robinson. Deter-
    mination of Pb complexation in oxic and sulfidic water
    using pseudovoltammetry. Environmental Science and
    Technology 37: 3845-3852 (2003).
                                                    19

-------
Ruby, M.V., A. Davis, and A. Nicholson.  In situ formation
    of lead phosphates in soils as a method to immobilize
    lead.   Environmental Science and Technology 28:
    646-654(1994).
Rudd, T., J.A. Campbell, and  J.N. Lester.  Use of model
    compounds to elucidate metal forms in sewage sludge.
    Environmental Pollution, 50: 225-242 (1988).
Saeki,  K., M. Okazaki, and S.  Matsumoto.  The chemi-
    cal phase changes in heavy metals with drying and
    oxidation of the lake sediments. Water Research, 27:
    1243-1251 (1993).
Scheckel, K.G., C.A. Impellitteri, J.A. Ryan, andT. Mcevoy.
    Assessment of a sequential extraction procedure for
    perturbed lead-contaminated samples with and without
    phosphorus amendments.  Environmental Science and
    Technology 37: 1892-1898 (2003).
Shannon, R.D. and J.R. White. The selectivity of a sequen-
    tial extraction procedure for the determination of iron
    oxyhydroxides and  iron sulfides in lake sediments.
    Biogeochemistry, 14: 193-208 (1991).
Shokes, T.E. and G. Moller.  Removal of dissolved heavy
    metals from  acid  rock drainage using  iron  metal.
    Environmental Science and Technology  33: 282-287
    (1999).
Sposito, G. The Surface Chemistry of Soils, Oxford Univer-
    sity Press, Oxford,  UK (1984).
Stumm, W. and K. Morgan. Aquatic Chemistry, 3rd Edition.
    Wiley-lnterscience, New York, NY (1996).
Sutherland, R.A. Comparison between non-residual Al, Co,
    Cu, Fe, Mn, Ni, Pb, and Zn released by  a three-step
    sequential extraction procedure and a dilute hydrochlo-
    ric acid leach for soil and road deposited  sediment.
    Applied Geochemistry 17:  353-365 (2002).
Tessier, A., P.G.C. Campbell, and M. Bison. Sequential ex-
    traction procedure for the speciation of particulate trace
    metals. Analytical Chemistry 51: 844-851 (1979).
Turekian, K.K. and K.H. Wedepohl.  Distribution of the ele-
    ments in some major units of the Earth's crust. Geologi-
    cal Society of America Bulletin 72: 175-192 (1961).
Trivedi, P., J.A. Dyer, and D.L. Sparks. Lead sorption onto
    ferrihydrite. 1. A macroscopic and spectroscopic as-
    sessment.   Environmental Science and Technology
    37:908-914(2003).
Uhler, A.D. and G.R. Helz. Solubility product  of galena at
    298K: A possible explanation for apparent supersatu-
    ration  in nature. Geochimica et Cosmochimica Acta
    48: 1155-1160(1984).
USEPA. Selection of Control Technologies for Remediation
    of Lead Battery Recycling Sites, EPA/540/2-91/014,
    U.S. Environmental  Protection Agency (1991).
USEPA. Report: Recent Developments for In Situ Treatment
    of Metals-contaminated Soils, U.S. Environmental Pro-
    tection Agency, Office of Solid Waste and Emergency
    Response (Draft) (1996).
USEPA.   Technology Alternatives  for the Remediation
    of Soils Contaminated with As, Cd, Cr, Hg, and Pb,
    EPA/540/S-97/500, U.S. Environmental  Protection
    Agency (1997).
USEPA. National Recommended Water Quality Criteria,
    4304T, Office of Water, Washington, DC. (2006).
Wang E.X., F.H. Bormann, and G. Benoit. Evidence of
    complete retention of atmospheric lead in the soils of
    northern hardwood forested ecosystems. Environmen-
    tal Science and Technology 29: 735-739 (1995).
Xue, H., L. Sigg, and F.G. Kari. Speciation of EDTA in natural
    water: Exchange  of FeEDTA in  river water.  Environ-
    mental Science and Technology 29: 59-68 (1995).
Yeates, W.S. and E.F Ayres. Plattnerite and its occurrence
    near Mullan, Idaho. American Journal of Science 43:
    407-412 (1892).
Zanker, H., H. Moll, W. Richter, V. Brendler, C. Hennig, T.
    Reich, A. Kluge, and G. Huttig. The colloid chemistry of
    acid rock drainage solution from an abandoned Zn-Pb-
    Ag mine. Applied Geochemistry 17: 633-648  (2002).
Zhang, P., J.A.  Ryan, and L.T. Bryndzia.  Pyromorphite
    formation from goethite adsorbed lead. Environmental
    Science and Technology 31: 2673-2678 (1997).
Zhang, P.  and J.A. Ryan.  Transformation of Pb(ll) from
    cerussite to chloropyromorphite in the presence of
    hydroxyapatite under varying conditions of pH. Environ-
    mental Science and Technology 33: 625-630 (1999).
                                                   20

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                                                 Nickel
                        Kirk G. Scheckel, Robert G. Ford, Richard T. Wilkin
Occurrence and Distribution

Industrial activity and natural environmental conditions have
led to the introduction of nickel into soil and aquatic environ-
ments as a result of anthropogenic and geogenic sources,
respectively (Duke, 1980; RichterandTheis, 1980). Nickel
is a relatively minor constituent of the earth's crust having
an average concentration  of less than 0.01% by weight
and  ranking 24th in terms  of abundance.  Nickel is very
heterogeneously distributed among crustal rocks  ranging
from less than 0.0001% in sandstone and granite to 4% in
coveted ore deposits (Duke,  1980).  Nickel can be found
in  igneous, sedimentary, and metamorphic rocks as well
as nickel ores.  In soils, nickel ranges from 5 - 500 mg kg-1
(Lindsay, 1979). Serpentine clay-rich  soils are noted for
natural geogenic abundance of nickel and have been the
focus for use of hyperaccumulating plants to phytomine
nickel (Chaney et al., 1995).

Nickel is one of the most mobile of the heavy metals in the
aquatic environment. The mobility of nickel in the aquatic
environment is controlled largely by competition between
various sorbents to scavenge it from solution and ligands to
form non-sorptive complexes. Although data are limited, it
appears that in pristine environments, hydrous oxides and
phyllosilicates control nickel  mobility via co-precipitation
and sorption. In polluted environments, the more preva-
lent organic compounds will keep nickel soluble by ligand
complexation.  In reducing environments, insoluble nickel
sulfide may form.  Nickel  chloride  is  water-soluble and
would be expected to release divalent nickel into the water.
The atmosphere is a major conduit for nickel as particulate
matter.  Contributions to atmospheric loading come from
both natural sources and anthropogenic activity, with input
from both stationary and mobile sources. Various  dry and
wet precipitation processes remove  particulate matter as
wash out or fallout from the atmosphere with transfer to
soils and waters.   Soil borne nickel  may enter waters by
surface runoff or by percolation into ground  water. Once
nickel is  in surface and ground-water systems,  physical
and chemical interactions (complexation, precipitation/dis-
solution,  adsorption/desorption, and oxidation/reduction)
occur that will determine its fate and that of its constituents.
The only gaseous  nickel compound of environmental im-
portance is nickel carbonyl. Under ambient  conditions in
moist air, it decomposes to form nickel carbonate. Thus,
in the atmosphere  at concentrations near the ppb level, it
has a half-life of about 30 minutes.  The removal of nickel
carbonyl  by precipitation or by adsorption on surfaces has
not been documented.  Since this compound is soluble
in  water, precipitation scavenging  is possible.  Nothing
is  known about  its  reaction with natural  surfaces or its
uptake by vegetation. Thus, dry deposition rates cannot
be predicted until some  experimental investigations have
been conducted.  Although nickel is bioaccumulated, the
concentration factors are such as to suggest that partition-
ing into the biota is not a dominant fate process.

Production of  nickel was  84.6 million  pounds  in 1986,
down slightly from 90 million pounds reported in  1982. In
1986 it was estimated that industries consumed nickel as
follows: transportation, 25%, chemical industry, 15%; elec-
trical equipment, 9%; construction, 9%; fabricated metal
products, 9%; petroleum, 8%; household appliances, 7%;
machinery, 7%; and other,  11%. Nickel carbonate is used
in nickel catalyst production for organic chemical manufac-
ture, petroleum refining and edible oil hardening.  Nickel
oxide consumption in 1972 (representing over 30  million
pounds containing nickel)  is estimated  to  have  been as
follows: 60% for stainless and heat resisting steels, 27%
for other steel alloys, 8% for other nickel alloys, 2% for cast
irons, and 3% for other uses (USEPA, 1986). From 1987
to  1993, according to the Toxics Release Inventory nickel
released to land and water totaled nearly 27 million pounds,
of which most was to land.  These releases were  primarily
from nickel smelting/refining and steelworks industries. The
largest releases occurred in Oregon and Arkansas.  The
largest direct releases to water occurred in Maryland and
Georgia (USEPA, 2003).

Plume Characteristics

The mobility of nickel in ground water will be controlled
by partitioning reactions to aquifer sediments.   Possible
mechanisms influencing nickel partitioning to subsurface
solids include direct adsorption to clay minerals, adsorp-
tion and/or coprecipitation with metal oxides, complexation
with natural organic particles, ion exchange with charged
surfaces, and direct precipitation as an hydroxide, carbon-
ate or sulfide (Snodgrass, 1980). The chemical speciation
of  nickel in solution exerts a significant influence on the
extent  and mechanism(s)  of partitioning to aquifer sedi-
ments, which may be influenced by acid-base reactions,
oxidation-reduction  reactions influencing  the  speciation
of  complexing inorganic solution species (e.g., aqueous
sulfate vs. sulfide), and interactions with dissolved organic
compounds. In general, inorganic/organic species that form
dissolved complexes with nickel tend to enhance transport
                                                    21

-------
of nickel in soil profiles to subsurface water (e.g., dissolved
organic carbon; Christensen et al., 1996; Warwick et al.,
1997; Christensen and Christensen, 2000; Friedly et al.,
2002). Field studies on transport in the subsurface illustrate
several general  conditions that are anticipated to result in
expanding  nickel  plumes, including 1) acidic conditions
(Kjoller et al., 2004), 2)  manganese-  and iron-reducing
conditions (Larsen and Postma,  1997), and 3) the pres-
ence of mobile organic compounds that form soluble nickel
complexes  (Christensen et al. 1996; Kent et al., 2002).

Remedial Technologies

Possible  engineered approaches that  can be employed
for remediation of a ground-water plume containing nickel
include physical  removal of contaminated soils or sediments
that serve  as a long-term source of nickel  leached into
ground water, extraction of the dissolved plume with some
method of above-ground treatment, physical isolation of the
dissolved plume, or in-situ treatment of a dissolved plume
resulting  in immobilization of dissolved nickel within the
aquifer. Of these technologies, the use of permeable reac-
tive barriers (PRBs) for the capture and immobilization  of
nickel plumes has been investigated and applied in field set-
tings due to favorable performance and cost characteristics
(Blowes et al., 2000). Both carbon- and  metallic iron-based
(or zero valent iron) reactive media have been employed for
nickel removal from ground water.  For carbon-based media,
nickel removal is generally considered to occur through the
precipitation of sulfide minerals, including nickel sulfides or
coprecipitation of nickel with iron sulfides (e.g., Ludwig et al.,
2002; McGregor et al., 2002). Reactive sulfide is generated
in this type of  PRB as a result of microbial sulfate reduction
stimulated by degradation of an organic carbon substrate
incorporated  into  the reactive barrier media.  Zerovalent
iron media have also been tested for the removal of nickel
in ground water (e.g., Wilkin and McNeil, 2003).  For this
material,  nickel removal may be achieved either through the
stimulation of sulfate reduction with precipitation  as a sulfide
or through coprecipitation with or adsorption onto metallic
iron corrosion products such as iron (hydr)oxides. There
is also laboratory and field evidence that nickel immobili-
zation can be enhanced through  the addition of chemical
amendments  that promote nickel precipitation within soil or
aquifer sediments (e.g., Lothenbach et al., 1997; Boisson
et al.,  1999; Seaman et al., 2001). The applicability and
performance of these technologies will depend on the geo-
chemical characteristics within the ground-water plume in
conjunction with the velocities of ground-water flow and the
flux of beneficial and non-beneficial reactive components
transported within the plume.

Regulatory Aspects

In 2005, nickel was ranked 55 of 275 hazardous substances
on the Comprehensive  Environmental  Response, Com-
pensation, and  Liability Act (CERCLA) National Priorities
List (NPL) based on frequency of  occurrence at NPL sites,
toxicity, and potential for human exposure to the substances
found at  NPL sites (ATSDR, 2005; http://www.atsdr.cdc.
gov/cercla/supportdocs/text.pdf).  There are  currently no
primary or secondary drinking water standards (maximum
contaminant level or MCL) in place for nickel in potable
water sources (USEPA, 2006a; See also http://www.epa.
gov/safewater/dwh/t-ioc/nickel.html). However, the health
advisory for nickel, an estimate of acceptable drinking water
levels for a chemical substance based on health effects in-
formation, via consumption of water has been set at 1 mg L~1
for one- and ten-day exposures for  a child and 0.1 mg L~1
for a lifetime exposure for an adult (USEPA, 2006a; http://
www.epa.gov/waterscience/criteria/drinking/dwstandards.
pdf).  For non-potable water sources, ambient water qual-
ity criteria (AWQC) that are protective of aquatic life may
serve as alternative cleanup goals.  For nickel, current
statutes list both acute and chronic criteria for fresh waters
as 0.47 mg L~1 and 0.052 mg L~1,  respectively, for a water
hardness of 100 mg L1  (USEPA, 2006b; http://www.epa.
gov/waterscience/criteria/nrwqc-2006.pdf).   Adjustments
to these criteria are to be applied for waters with different
hardness. An example of where this criterion may apply
is a site where contaminated ground water discharges to
surface water.

Geochemistry and Attenuation Processes

Aqueous Speciation

In ambient aqueous systems, nickel exists in the divalent
oxidation state and is not subject to oxidation-state trans-
formations under typical conditions.  Nickel predominantly
exists as a cationic species (Ni2+) or various hydolysis spe-
cies (e.g., NiOH+) at near-neutral pH (Baes and  Mesmer,
1986). However, nickel may also form dissolved complexes
in the presence of high  concentrations of inorganic ions
such as carbonate/bicarbonate and sulfate (Hummel and
Curti, 2003; Chen et al., 2005) or organic  ligands such
as natural/synthetic carboxylic acids and dissolved humic
compounds (Bryce and Clark, 1996; Baeyens et al., 2003;
Strathman and Myneni, 2004).  It  is anticipated that nickel
may form complexes with dissolved sulfide under sulfate-
reducing conditions, although the current state of knowledge
is insufficient to ascertain the relative importance of these
species in aqueous systems (Thoenen, 1999). The forma-
tion of solution complexes, especially with organic ligands,
may limit sorption of nickel to mineral surfaces in aquifer
sediments (see Adsorption section  below).

Solubility

Nickel may be immobilized within  ground water through
formation of pure nickel  precipitates such as hydroxides,
silicates, or sulfides (Merlen et al.,  1995; Mattigod et al.,
1997;Scheideggeretal., 1997; Thoenen, 1999; Scheinost
and Sparks, 2000; Peltier et al., 2006) or through copre-
cipitation with other soil forming minerals such as silicates,
iron oxides/sulfides, or carbonates (Manceau et al., 1985;
Manceau and Galas, 1986; Huerta-Diaz and Morse, 1992;
Ford et al., 1999a; Hoffmann and Stipp, 2001). Predicted
nickel concentrations in the absence of sulfide for several
potential pure  nickel precipitates are shown in Figure 3.1.
These data suggest that phyllosilicate and layered double
hydroxide (LDH) precipitates (incorporating aluminum) may
result in dissolved nickel concentrations below most relevant
regulatory criteria over a pH range typical for ground water.
                                                     22

-------
These data also point to the limited capability of pure nickel
carbonates and hydroxides in controlling dissolved nickel
concentrations to sufficiently low values except under very
alkaline conditions. In the  presence of dissolved sulfide,
the precipitation of a nickel sulfide may plausibly control the
concentration of dissolved  nickel.  The Eh-pH conditions
under which these solubility-limiting  phases may form  is
shown in Figure 3.2. According to these data, nickel-bear-
ing phyllosilicate and/or LDH precipitates possess  large
stability fields  indicating their relative importance to con-
trolling nickel solubility under a range of conditions. These
calculations point to the importance of dissolved aluminum
and silicon concentrations in ground water relative to the
potential sequestration of nickel via precipitation (Ford et al.,
1999b; Scheinost et al., 1999). As discussed below (see
Adsorption section), the formation of these nickel-bearing
precipitates may be facilitated through initial  adsorption
onto clay minerals within the aquifer.
                                                                  Ni phyllosilicate
                                                                  Ni-AI-SO4 LDH
                                                                  Ni-AI-CO3 LDH
                                                                          b)
                      -20

                      --24,

                      --28.

                      --32
                                                                     I  ' I ' I
                                                                     8  9 10 11
Health Advisory
(Lifetime Exposure)

\ PH
AWQC
(Chronic, Freshwater, 100 mg/L Hardness)
                                                                    PH
Figure 3.1  (a) Predicted solubility of various Ni precipitates that could form in aerobic ground water with concentra-
            tions of Al and Si controlled by the solubility of the clay mineral, kaolinite. (Note y-axis break to expand
            lower end of scale.) (b) Expansion of dissolved Ni data for equilibrium with Ni phyllosilicate; plotted on
            logarithmic scale. Nominal ground-water composition: 0.005 mole/L NaCI, 0.001 mole/L K2SO4, 0.001
            mole/L MgNO3, 0.001 mole/L CaCO3, and 34 u mole/L Ni (2000 \jg Ni/L). Model predictions using Visual
            MINTEQ Version 2.50 (Based on MINTEQA2 described in Allison et al. (1991); available at http://www.
            lwr.kth.se/English/OurSoftware/vminteq/) with solubility constants added for Ni phyllosilicate, Ni-AI-SO4
            LDH,  and Ni-AI-CO3 LDH (Peltier et al., 2006); total dissolved Ni concentrations modeled individually for
            each Ni solid phase by setting kaolinite as an  'infinite' solid and the Ni solid phase as a 'possible' solid for
            each pH titration.
                   a)
b)
                                                        w
                                    pH
                  PH
Figure 3.2 Eh-pH diagrams for nickel at 25 C. (a) System Ni-H2O-Ca-AI-NO3-HCO3-SO4 (2 mg Ni/L; 40 mg Ca/L;
            3 mg AI/L; 6 mg NO/L; 60 mg HCO/L; 100 mg SO/L).  Stability fields for solids are shaded green
            (Vaesite = NiS2). (b) Same system plus 3 mg Si/L.  Thermodynamic data for Ni3Si4O10(OH)2 and
            Nioe3AI033(OH)2(SO4)0125 are from Peltier et al. (2006). [Note that the solubility of the Ni-AI-SO4 LDH was
            adjusted to correct for charge imbalance for the chemical structure published in Peltier et al. (2006).]
                                                      23

-------
Attenuation  of  nickel may also occur via coprecipita-
tion during the formation  of (hydr)oxides or sulfides of
iron. These minerals have been observed to form at the
boundaries between oxidizing and reducing zones within
ground-water plumes. There are numerous laboratory and
field observations that demonstrate the capacity of these
precipitates for nickel uptake (Schultz et al., 1987; Huarta-
Diaz and Morse, 1992; Coughlin and Stone, 1995;  Ford et
al., 1997; Ford etal., 1999a).  Under these circumstances,
the solubility of nickel will depend on the stability of the host
precipitate phase.  For example,  iron oxide  precipitates
may alternatively transform to more stable forms (Ford et
al., 1997), stabilizing coprecipitated nickel over the long
term, or these precipitates may dissolve concurrent with
changes  in ground-water redox chemistry (e.g., Zachara
etal.,2001).

Adsorption

Adsorption of nickel in soil environments is dependent on
pH, temperature, and type of sorbent (minerals or organic
matter), as well as the concentration of aqueous  complex-
ing agents, competition from other adsorbing cations, and
the ionic strength in ground water.  Nickel has been shown
to adsorb onto many solid components encountered in
aquifer sediments, including iron/manganese oxides, clay
minerals (Dahn et al., 2003; Bradbury and Baeyens, 2005),
and solid organic matter (Nachtegaal  and Sparks, 2003).
Sorption to iron/manganese oxides and clay minerals has
been shown to be of particular importance for controlling
                   nickel mobility in subsurface systems. The relative affinity
                   of these individual minerals for nickel uptake will depend
                   on the mass distribution of the sorbent minerals as well
                   as the predominant geochemical conditions  (e.g., pH and
                   nickel aqueous speciation).  For example, the pH-depen-
                   dent  distribution of nickel between iron and manganese
                   oxides [hydrous ferric oxide (HFO) and a birnessite-like
                   mineral (nominally MnO2)] for a  representative  ground-
                   water composition is shown in Figure 3.3a.  Based on the
                   available compilations for surface complexation constants
                   onto  these two solid phases (Dzombak and Morel, 1990;
                   Tonkin et al., 2004), one would project the predominance
                   of nickel sorption to MnO2 at more acidic pH and the pre-
                   dominance of HFO (or ferrihydrite) at more basic pH. With
                   increasing  mass  of MnO2, the solid-phase  speciation  of
                   nickel will be progressively dominated by sorption to this
                   phase.  There are examples of the relative  preference  of
                   nickel sorption to manganese oxides over iron oxides for
                   natural systems (e.g., Larsen and Postma, 1997; Manceau
                   et al., 2002; Kjoller et al., 2004; Manceau et al., 2006). As
                   shown in Figure 3.3b, nickel adsorption may be inhibited
                   (or nickel desorption enhanced) through the formation  of
                   solution  complexes with organic  ligands  such as EDTA
                   or  natural organic matter (e.g., Bryce  and Clark, 1996;
                   Nowacketal., 1997). These dissolved compounds may be
                   present as natural components within ground water or as
                   co-contaminants within a contaminant plume (e.g., Means
                   etal., 1978).
                     2000-
                     1600-
                     1200-
                     800-
                     400-
                       0-
                                                     a)  -
           [Ni]
       --[Ni]
        A-  [Ni]
        *  [Ni]
          aq

          sorted, total  _

          sorted, HFO

          sorted, MnO2
                                             Health Advisory
                                           (Lifetime Exposure)
                                \
                                5
\
6
\
7
\
8
 \  '\
10  11
\
5
\
6
\  '\  '\
9   10   11
                                         PH
                                     PH
Figure 3.3  (a) Nickel sorption as a function of pH in the presence of an hypothetical aquifer sediment with iron and
            manganese oxides reflective of the crustal abundance of these elements (Schulze, 2002; assumed 30%
            porosity with 185.0 g HFO/L and 1.66 g MnO/L).  (b) Same conditions as in (a), but with 10 u/W EDTA
            added. Nominal ground-water composition: 0.005 mole/L NaCI, 0.001 mole/L K2SO4, 0.001 mole/L
            MgNO3, 0.001 mole/L CaCO3, and 34 u mole/L Ni (2 mg Ni/L). Model predictions using Visual MINTEQ
            Version 2.50 (Based on MINTEQA2 described in Allison et al. (1991); available at http://www.lwr.kth.
            se/English/OurSoftware/vminteq/) with available surface complexation parameters derived from Dzombak
            and Morel (1990) and Tonkin et al. (2004); kaolinite set as an 'infinite'solid forpH titration.
                                                    24

-------
Table 3.1  Natural attenuation and mobilization pathways for nickel.
   Attenuation Processes
 Mobilization Processes
         Characterization Approach
 Precipitation of Ni as carbonate,
 layered double  hydroxide,  or
 phyllosilicate phase in oxidized/
 reduced systems; precipitation of
 Ni as a sulfide in sulfate-reducing
 systems
Dissolution of  Ni precipitates
due  to  decreased  pH; dis-
solution of NiS due to shift
from  reducing to oxidizing
conditions.
Evaluation of dissolved Ni concentration in ground
water.  Determination of total Ni in the solid matrix
and suspected components in Ni-bearing precipitate.
Evaluation of mineral solubility relative to ground-wa-
ter chemistry and published solubility constants.
 Co-precipitation of Ni as a trace
 component in oxyhydroxides or
 sulfides of iron or manganese
Dissolution of host oxyhydrox-
ide due to decrease in pH or
shift from oxidizing to reducing
conditions; dissolution of host
sulfide due to shift from reduc-
ing to oxidizing conditions.
Evaluation of Ni concentration in ground water and in
solid matrix. Evaluation of host precipitate formation
relative to existing ground-water chemistry; determi-
nation of host mineral content in aquifer sediments
via mineralogical characterization. Evaluation of Ni
solid-phase partitioning using sequential extraction
methodologies.
 Adsorption of Ni to  iron oxyhy-
 droxides, iron sulfides, or other
 mineral surfaces
Desorption due to low pH, high
competing cation concentra-
tions, or high DOC concentra-
tions for oxyhydroxides and
sulfides. Reductive dissolution
of iron hydroxides or oxidative
dissolution of iron sulfides.
Evaluation of Ni concentration in ground water and in
solid matrix. Evaluation of Ni solid-phase partitioning
using sequential extraction methodologies.  Batch
and column testing to determine Ni uptake behavior
and capacity of site-specific aquifer materials under
variable geochemical conditions.
As  previously noted,  adsorption of nickel onto  mineral
surfaces may serve as a precursory step to the formation
of trace precipitates that reduce the potential for desorp-
tion with changes  in ground-water chemistry.  This may
be realized through the nucleation and growth of  surface
precipitates  on clay  mineral surfaces due to continued
uptake of nickel (Scheckel and Sparks, 2000; Scheckel et
al., 2000; Scheckel and Sparks, 2001; Dahn et al., 2002).
This type of process may compete with other adsorption
processes, such as ion exchange, depending on the pre-
vailing ground-water chemistry and characteristics of the
clay mineral (Elzinga and Sparks, 2001).

Site Characterization

Overview

Nickel mobility in ground water  is governed  by the total
concentration of nickel, the distribution  of nickel species
in water, and the nature of nickel partitioning in the solid
phase.  The development  of site-conceptual models  for
predicting the long-term fate of nickel at a contaminated
site will require information on the distribution and con-
centration of nickel in the  aqueous phase and the solid
phase.  Table 3.1  indicates possible natural attenuation
and mobilization pathways for nickel. Details  of the types
of analytical measurements  that may be conducted  on
sampled ground water and aquifer sediments to assist in
identifying the attenuation mechanism(s) are discussed in
the following paragraphs.
                       Aqueous Measurements

                       The total concentration of nickel in aqueous samples can
                       be determined by an array of methods ranging significantly
                       in sensitivity, detection limits, and accuracy. For aqueous
                       systems,  nickel can be  measured by flame/graphite fur-
                       nace atomic absorption (FAAS and GFAAS, respectively),
                       inductively coupled plasma atomic emission spectrometry
                       (ICP-AES) or  mass spectrometry (ICP-MS), colorimetry,
                       ion chromatography, and electrochemical methods (Stoep-
                       pler, 1980).  The standard colorimetric method for nickel
                       is the dimethylglyoxime (DMG) method (Amacher, 1996).
                       Ion chromatography works  well for nickel  in determining
                       total nickel in soil digestion solutions (Basta and Tabatabai,
                       1990).  Electrochemical  methods (e.g., anodic stripping
                       voltammetry, platinum electrode differential oscillopolarog-
                       raphy, or differential pulse polarography) are well suited for
                       aqueous samples and often employ DMG-coated electrodes
                       to concentrate nickel for better sensitivity (Stoeppler, 1980).
                       Of the list above, FAAS, GFAAS, ICP-OES, or ICP-MS are
                       the most common methods employed.

                       For ground water  with elevated  concentrations  of
                       dissolved organic carbon  (e.g., landfill  leachates)  or
                       known organic co-contaminants such  as EDTA, it  may
                       be necessary to determine the chemical  speciation  of
                       dissolved nickel.  Geochemical speciation  models  (e.g.,
                       MINTEQA2, PHREEQC, EQ3/6) may be employed  to
                       assist in determining  aqueous nickel speciation, but the
                                                    25

-------
accuracy of existing codes and/or associated geochemical
databases  for  assessing complexation  with  dissolved
organic carbon has been questioned (e.g., Christensen
and Christensen, 2000). These computer speciation codes
also require, at a minimum, the  concentrations of major
anions, major cations, total organic  carbon (or specific
species of  organic compounds), temperature,  and pH.
Direct determination of the fraction of organic-complexed
nickel may be accomplished through analytical fractionation
of nickel using various exchange resins (e.g., Christensen
and Christensen, 2000; Jian and Presley, 2002).  As noted
by Jiann and  Presley (2002), the approach  to sample
preservation prior to separation may depend on whether
fractionation can be conducted in the field or at a later time
in the laboratory. Voltammetric measurements may also be
employed to examine the relative distribution and stability
of nickel complexes with dissolved organic compounds in
water (e.g., Van den Berg and Nimmo, 1987; Bedsworth and
Sedlak, 1999).  Supporting data for these measurements
would  include determinations of total dissolved organic
carbon along with specific organic constituents suspected
in the ground-water plume.  In addition, while regulatory
requirements stipulate  that unfiltered  ground-water
samples  be analyzed  to  support regulatory decisions  at
a contaminated site, it may be necessary to also collect
filtered samples to help define the process(es) controlling
contaminant mobility. The use of  0.45 urn  pore size filters
is common  as an arbitrary cutoff point  to differentiate
between dissolved and particulate  phases   in water
samples. However, caution is recommended when using
this approach, particularly for Fe and Al and other elements
that may be associated with Fe or Al particles (including Ni)
that could pass through 0.45 urn  filters.  The use of filters
with pore sizes less than 0.1  urn will generally provide a
better assessment of the dissolved vs. particulate load in
ground water.

It has also been observed that nickel may be leached from
certain grades of stainless steel well casing/screen materi-
als under chemical conditions that may be encountered in
contaminant plumes. Two published studies have provided
detailed evaluation of the extent of nickel (and chromium)
leaching that may occur for type 304 or 316 stainless steel
screens (Hewitt, 1994; Oakley and Korte, 1996). Oakley
and  Korte (1996) provide a site-specific example of  how
elevated  nickel concentrations derived from continuous
leaching of  well screen materials  may be falsely  identified
as a component of a ground-water plume. This suggests
that careful consideration should  be given to the types  of
well  screen  materials  and sampling protocols  employed
relative to the assessment of potential contaminants  of
concern within a plume.

Solid Phase Measurements

The  implementation of an analytical approach to identify
nickel speciation in aquifer sediments is a challenging pro-
cess. The accuracy of the analytical finding  is dependent on
the method of sample collection/preservation and the tools
used to identify the mechanism of nickel partitioning.  It is
recommended that the analytical protocol be designed  to
address the potential redox sensitivity of the solid phase(s)
to which nickel may be partitioned (e.g., sulfides in reduced
sediments). Tools to  evaluate the mechanism of nickel
solid phase partitioning range in complexity from relatively
simple chemical extractions to advanced spectroscopic
techniques.

Bulk solid phase nickel concentration can be determined
directly on the solid sample by X-ray fluorescence spec-
trometry, neutron activation analysis, or following chemical
digestion and analysis of nickel in the resultant liquid phase.
A variety of digestion or extraction methods can be found in
the literature (Amacher, 1996). Neutron activation analysis
is not commonly employed due to the scarcity of neutron
sources required to irradiate the sample. X-ray fluorescence
is the most attractive approach due to the relative ease
of sample preparation, which may be conducted with the
sample in its original state or following fusion with lithium
metaborate.  When combined  with the determination of
other major or trace elements in the solid sample, this
provides an initial step for assessing possible association
of nickel with various solid phase components.  This  type
of analysis can be conducted on the bulk sample as well
as at a  microscopic level using wavelength (electron mi-
croprobe) or energy dispersive spectroscopy coupled to a
scanning or transmission electron microscope. Microscopic
examination allows one to better differentiate whether nickel
may be distributed across a number of different mineral
phases within the solid sample or primarily associated with
a discrete phase. There are limitations to this approach
(Pye, 2004), a significant one being that the analysis does
not necessarily provide unique  mineral identification ne-
cessitating the collection of supporting mineralogical and
chemical data.

More detailed  information on  the specific partitioning
mechanism(s) controlling  nickel solid phase speciation is
typically required to adequately support site assessment
for potential reliance on natural attenuation as part of a site
remedy. There have been many applications of sequential
extraction schemes to assess the speciation of solid phase
nickel (e.g., Tessier et al., 1979; Ryan et al., 2002; Peltier et
al., 2005; Buanuam et al., 2006). As discussed in the cited
reports, sequential extraction methods provide a useful tool
to assist in determining the  chemical  speciation of trace
metals in soils/sediments, but essentially all documented
methods show analytical limitations in selectively extract-
ing nickel and other metals associated with specific solid
components. Where feasible, it is recommended that com-
plimentary analytical techniques be employed to confirm the
accuracy of nickel speciation (e.g., D'Amore et al., 2005;
Manceau et al., 2006) or the accuracy of the extraction of
a targeted phase(es) for a given extractant (e.g., Shannon
and White, 1991; Ryan etal., 2002; Peltier etal., 2005). As
an example, Peltier et al. (2005)  have demonstrated that
a common extraction method employed to target metals
associated with easily reducible iron (hydr)oxides may also
dissolve iron sulfides that may be present. The results from
this analysis may lead to misidentification of a nickel asso-
ciation with iron (hydr)oxides, resulting in the development of
a conceptual site model that misrepresents the site-specific
                                                     26

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attenuation process.  Under reducing conditions, it is also
critical that aquifer sediments be sampled and processed
in a manner that prevents exposure to oxygen prior to ex-
traction in order to limit oxidation of reduced minerals (e.g.,
iron sulfides) that may host nickel.  Determination of the
host mineral phase(es) dissolved for each extraction step
is recommended, along with the use of surrogate Ni-bear-
ing phases spiked into the sediment to confirm accuracy
of the procedure (e.g.,  Rudd et al., 1988).  The choice of
appropriate nickel surrogate phases would be governed by
site-specific geochemical conditions or characterization of
the mineralogy of the aquifer sediment.

Long-term Stability and Capacity

The stability of attenuated nickel will largely depend on the
stability of  site-specific  geochemical conditions through
time.  For example, if nickel attenuation follows a pathway
of  coprecipitation with  iron  sulfide,  then  the  long-term
stability of  attenuated nickel will depend, in part,  on the
persistence of reducing conditions. If ground-water redox
conditions were to shift to oxidizing conditions, nickel might
be expected to release from the solid phase. It is therefore
important to understand the attenuation mechanism(s) so
that geochemical triggers for mobilization can be anticipated
and incorporated into evaluations of long-term monitoring
data.  For any proposed and identified attenuation mecha-
nism, there will exist possible scenarios whereby remobi-
lization could occur (i.e., changes in pH or Eh).  It will be
essential to explore the likelihood of such changes in site
geochemistry and the sensitivity of the attenuation pathway
to changes in the prevailing geochemical conditions.

Quantifying the attenuation capacity (as defined in Volume
1) will also require an understanding of the specific attenu-
ation pathway(s). Attenuation capacity, for example, could
be related to the extent that pH is buffered,  the availability
of sorptive sites in aquifer materials, or to  the supply of
electron donors needed to sustain microbially mediated re-
dox conditions.  For any proposed attenuation mechanism,
there will be assumptions  built into capacity estimations,
so it is recommended that uncertainty analysis accompany
capacity calculations.

Tiered Analysis
Determination of the viability of nickel remediation in ground
water via monitored natural attenuation will depend upon
proper assessment of contaminant loading  to the aquifer
and  prevailing geochemistry  and mineralogy within the
contaminant plume and the down gradient zone prior to
the point(s) of compliance. The goal of site  assessment is
to demonstrate the process(es) controlling  nickel seques-
tration onto aquifer solids and the  long-term stability of
solid phase nickel as a function of existing and anticipated
ground-water chemistry.  The  following tiered analysis
structure for site characterization provides an approach to
evaluate candidate sites and define the potential limitations
of MNA as part of a remedy for ground-water cleanup.

Tier I- Site characterization under Tier I will involve demon-
stration that the plume is static or shrinking, has not reached
compliance boundaries, and does not impact existing water
supplies.  Once this is established through ground-water
characterization, evidence is collected to demonstrate Ni
partitioning to aquifer solids within the plume.  If natural
attenuation processes are  active throughout the plume,
then there should be an observed increase in solid phase
concentrations within  regions  of  the  plume with higher
aqueous concentrations, e.g., near the source term.  This
field partitioning data may be supplemented by geochemical
modeling that incorporates measured water chemistry (e.g.,
pH, Eh, and major ion chemistry) throughout the plume to
assess the potential for solubility control by a nickel pre-
cipitate such as phyllosilicate or sulfide phase.  Identifica-
tion of active sequestration to  prevent nickel migration in
ground-water provides justification for proceeding to Tier
II characterization efforts.

Tier II -  Under Tier II, the apparent rate and mechanism(s)
of attenuation are determined.  Estimates of a site attenu-
ation  rate(s) can  be assessed via a well transect along
the ground-water flow path.  In addition, time-series  data
may be collected  at one or  more monitoring  points within
the plume. This information will allow assessment of the
relative timescales for contaminant immobilization and fluid
transport and determination of whether remediation objec-
tives can be met within the required regulatory time frame.
In addition, the mechanism(s)  for attenuation need to be
identified under this stage of site characterization.  This
effort may require determination of the chemical speciation
of aqueous and solid phase Ni, which may be approached
according to the following scheme:

 1)   Determination of  nickel solution speciation via di-
     rect analytical measurements in combination with
     speciation calculations based on characterized
     ground-water chemistry;
 2)   Calculation of the saturation state of ground water
     relative to measured aqueous chemistry compli-
     mented by the possible  isolation of discrete Ni
     mineral phases via density separations (or other
     schemes) in regions of the aquifer with highest
     solid phase concentrations;
 3)   Determination of  aquifer mineralogy to determine
     the relative abundance of  components with docu-
     mented capacity  for Ni sorption (e.g., Amonette,
     2002);
 4)   Identification of  nickel association(s)  with the
     various solid phase components of aquifer solids
     through combination of chemical extractions with
     microscopic/spectroscopic confirmation of phase
     associations, and;
 5)   Demonstration  of concurrence between the site
     conceptual model and mathematical model(s) that
     describe nickel removal mechanism(s).
It is recommended that identification of nickel chemical spe-
ciation in aqueous and solid matrices be conducted using
samples collected in a manner that preserves the in-situ
distribution of dissolved nickel and mineralogy and prevents
loss of nickel from aqueous samples (e.g., due to oxidation
and precipitation of ferrous iron in anoxic ground water).
                                                     27

-------
The demonstration of concurrence between conceptual and
mathematical models describing nickel transport will entail
development of site-specific parameterization of the chemi-
cal processes controlling nickel solid phase partitioning.

Tier III - Once the partitioning mechanism(s)  have been
identified for the site, the subsequent characterization ef-
fort under Tier III will involve determination of the stability
of immobilized Ni and the capacity of the aquifer to sustain
continued uptake.  It is recommended that the stability of
immobilized Ni be tested based on the anticipated evolution
of ground-water chemistry concurrent with plume shrinkage.
For example,  changes in ground-water pH can exert a sig-
nificant influence on Ni adsorption or precipitate solubility.
Therefore, it is recommended that sediment leach tests be
conducted to characterize the magnitude of Ni mobilization
as a function of pH for a ground-water chemistry representa-
tive of site conditions. It is recommended that the capacity
for Ni uptake  onto aquifer solids be  determined relative to
the specific mechanism(s) identified  in Tier II.  For example,
if site characterization under Tier II indicated that precipita-
tion of Ni sulfide due to microbial degradation of organic
compounds coupled with sulfate reduction occurs within the
aquifer, then it is recommended that the mass distribution
of organic carbon and sulfate to support this reaction within
the aquifer be determined. This site-specific capacity can
then be compared to Ni mass loading within the  plume in
order to assess the longevity of the natural attenuation
process.  If site-specific tests  demonstrate the stability of
immobilized Ni and sufficient capacity within the aquifer to
sustain Ni attenuation, then the site  characterization effort
can progress to Tier IV. For cases where contaminant stabil-
ity is sufficient but aquifer capacity is insufficient for capture
of the entire plume, then a determination of the benefits of
contaminant source reduction may be necessary.

Tier IV- Finally, under Tier IV a monitoring plan  is estab-
lished along with contingency plans in the event of MNA
failure.  It is  recommended that the  monitoring plan be
designed to establish both continued plume stability and to
identify changes in ground-water chemistry that may lead
to re-mobilization of attenuated Ni.  The specific chemical
parameters to be monitored will include those identified
under Tier III  that may  halt Ni partitioning to aquifer sedi-
ments and/or  result in solubilization of either discrete Ni pre-
cipitates or aquifer minerals that sequester Ni from ground
water. For example, solution phase  parameters that could
alter either Ni precipitation or adsorption include increases
in soluble organic carbon in combination with changes in
ground-water pH. In contrast, the concentration of dissolved
iron or sulfate may indicate the dissolution of an important
sorptive phase within the aquifer (e.g., reductive dissolution
of iron oxides or oxidative dissolution of su If ides). Changes
in these parameters may occur prior to observed changes
in solution Ni and, thus, serve as monitoring triggers for
potential MNA failure.  In this instance, a contingency plan
can be implemented that incorporates strategies to arrest
possible plume expansion beyond compliance boundaries.
Possible  strategies  to  prevent plume expansion include
pump and treat operations, installation of reactive barriers
to enhance uptake capacity perpendicular to the direction
of plume advance, or enhancement of natural attenuation
processes within the aquifer through the injection of soluble
reactive components.

References

Allison,  J.D.,  D.S.  Brown,  and K.J. Novo-Gradac. MINT-
    EQA2/PRODEFA2, A Geochemical Assessment Model
    for Environmental Systems: Version 3.0 User's Manual.
    United States  Environmental Protection Agency, Of-
    fice of Research and Development, Washington, DC,
    EPA/600/3-91/021 (1991).
Amacher, M.C. Nickel, cadmium, and lead. In Methods of
    Soil Analysis: Chemical Methods, Part 3. D.L. Sparks,
    (Ed.), Soil Science Society of America, Madison, Wl,
    pp. 739-768(1996).
Amonette, J.E. Methods for determination of mineralogy and
    environmental  availability. In So/7 Mineralogy with En-
    vironmental Applications. J.B. Dixon and D.G. Schulze
    (Eds.), Soil Science  Society of America Book Series.
    Soil Science Society of America, Inc., Madison, Wl,
    pp. 153-197(2002).
ATSDR  CERCLA Priority List of Hazardous Substances
    that will be the Subject of Toxicological Profiles and
    Support Document, Department of Health and Human
    Services (2005).
Baes, C.F and R.E. Mesmer.  The Hydrolysis of Cations.
    Malabar, FL, Krieger Publishing Company (1986).
Baeyens, B., M.H.  Bradbury, and W. Hummel. Determina-
    tion of aqueous nickel-carbonate and nickel-oxalate
    complexation constants. Journal of Solution Chemistry
    32:319-339 (2003).
Basta, NT. and M.A.Tabatabai. Ion chromatography deter-
    mination of total metals in soils. So/7 Science Society
    of America Journal 54: 1289-1297 (1990).
Bedsworth, W.W. and D.L. Sedlak. Sources and environmen-
    tal fate of strongly complexed nickel in estuarine waters:
    The role of ethylenediaminetetraacetate. Environmental
    Science and Technology 33: 926-931 (1999).
Blowes, D.W., C.J. Ptacek, S.G. Benner, C.W.T McRae, T.A.
    Bennett, and  R.W. Puls. Treatment  of inorganic con-
    taminants using permeable reactive barriers.  Journal
    of Contaminant Hydrology 45: 123-137 (2000).
Boisson, J., M. Mench, J. Vangronsveld, A. Ruttens, P. Kop-
    ponen, and T. De Koe. Immobilization of trace metals
    and arsenic by different soil additives: Evaluation by
    means of chemical extractions. Communications in Soil
    Science and Plant Analysis 30: 365-387 (1999).
Bradbury, M.H.  and B.  Baeyens. Modelling the sorption
    of Mn(ll), Co(ll), Ni(ll), Zn(ll), Cd(ll), Eu(lll),  Am(lll),
    Sn(IV), Th(IV),  Np(V)  and U(VI) on montmorillonite:
    Linear free energy relationships and estimates of sur-
    face binding constants for some selected heavy metals
    and actinides. Geochimica et Cosmochimica Acta 69:
    875-892 (2005).
Bryce, A.L. and S.B. Clark. Nickel desorption kinetics from
                                                     28

-------
    hydrous ferric oxide in the presence of EDTA. Colloids
    and Surfaces A - Physicochemical and Engineering
    Aspects 107: 123-130 (1996).
Buanuam, J.,  K. Tiptanasup, J. Shiowatana, M. Miro, and
    E.H. Hansenc. Development of a simple extraction cell
    with bi-directional continuous flow coupled on-line to
    ICP-MS for assessment of elemental associations in
    solid samples. Journal of Environmental Monitoring 8:
    1248-1254(2006).
Chaney, R., S. Brown, Y. M. Li, J.S. Angle, F. Homer, and
    C. Green. Potential use of metal hyperaccumulators.
    Mining Environmental Management September: 9-11
    (1995).
Chen, T., G. Hefter, and R. Buchner.  Ion association and
    hydration in aqueous solutions of nickel(ll) and cobalt(ll)
    sulfate. Journal of Solution Chemistry 34: 1045-1066
    (2005).
Christensen, J.B. and T.H. Christensen. The effect  of pH
    on the complexation of Cd,  Ni and Zn by dissolved
    organic carbon from  leachate-polluted  groundwater.
    Water Research 34: 3743-3754 (2000).
Christensen, J.B., D.L. Jensen, and T.H. Christensen. Effect
    of dissolved organic carbon on the mobility of cadmium,
    nickel and zinc in leachate polluted groundwater.  Water
    Research 30: 3037-3049 (1996).
Coughlin, B.R. and AT. Stone. Nonreversible adsorption of
    divalent ions (Mnll, Coll, Mill, Cull, and Pbll) onto goe-
    thite: effects of acidification, Fell addition and picolinic
    acid addition. Environmental Science and Technology
    29:2445-2455(1995).
Dahn, R. A.M. Scheidegger, A. Manceau, M.L. Schlegel, B.
    Baeyens, M.H. Bradbury, and M. Morales. Neoformation
    of Ni phyllosilicate upon Ni uptake on montmorillonite: A
    kinetics study by powder and polarized extended X-ray
    absorption fine structure spectroscopy. Geochimica et
    Cosmochimica Acta 66: 2335-2347 (2002).
Dahn, R.,  A.M. Scheidegger, A. Manceau, M.L. Schlegel,
    B. Baeyens, M.H. Bradbury, and D. Chateigner. Struc-
    tural evidence for the sorption of Ni(ll) atoms on the
    edges of  montmorillonite clay minerals: a polarized
    X-ray absorption fine structure study. Geochimica et
    Cosmochimica Acta 67: 1-15 (2003).
D'Amore, J.J., S.R. AI-Abed, K.G. Scheckel, and J.A. Ryan.
    Methods for speciation of metals in soils: A review. Jour-
    nal of Environmental Quality 34:1707-1745 (2005).
Duke, J.M. Nickel in rocks and ores. In Nickel in the Envi-
    ronment. J.O. Nriagu (Ed.), New York, NY, John Wiley
    &Sons, pp. 27-50(1980).
Dzombak, D.A. and F.M.M. Morel. Surface Complexation
    Modeling: Hydrous Ferric Oxide.  John Wiley &  Sons,
    New York, NY (1990).
Elzinga, E.J. and D.L.  Sparks. Reaction condition effects
    on nickel sorption mechanisms in illite-water suspen-
    sions.  So/7 Science Society of America Journal 65:
    94-101 (2001).
Ford, R.G., P.M. Bertsch, and K.J. Farley. Changes in tran-
    sition and heavy metal partitioning during hydrous iron
    oxide aging. Environmental Science and Technology
    31:2028-2033(1997).
Ford, R.G., K.M. Kemner, and P.M. Bertsch. Influence of
    sorbate-sorbent interactions on the crystallization kinet-
    ics of nickel- and lead-ferrihydrite coprecipitates. Geo-
    chimica et Cosmochimica Acta 63: 39-48 (1999a).
Ford, R.G., A.C. Scheinost, K.G. Scheckel, and D.L. Sparks.
    The link between clay  mineral weathering and  the
    stabilization of Ni surface precipitates. Environmental
    Science and Technology 33: 3140-3144 (1999b).
Friedly, J.C., D.B. Kent, and  J.A. Davis.  Simulation of the
    mobility of  metal-EDTA complexes in groundwater:
    The influence of contaminant metals. Environmental
    Science and Technology 36 3: 355-363 (2002).
Hewitt, A.D. Dynamic study of common well screen ma-
    terials. Ground Water Monitoring and Remediation 9:
    87-94(1994).
Hoffmann, U. and S.L.S. Stipp. The behavior of Ni2+ on
    calcite surfaces. Geochimica  et Cosmochimica Acta
    65:4131-4139 (2001).
Huerta-Diaz, M.A. and  J.W. Morse. Pyritization of trace
    metals  in anoxic marine sediments. Geochimica et
    Cosmochimica Acta 56:  2681-2702 (1992).
Hummel, W. and E.  Curti. Nickel aqueous speciation and
    solubility at ambient conditions: a thermodynamic elegy.
    Monatshefte furChemie 134: 941-973 (2003).
Jiann, K.-T and B.J. Presley. Preservation and determina-
    tion of trace metal partitioning in river water by a two-
    column ion exchange method. Analytical Chemistry74:
    4716-4724(2002).
Kent, D.B., J.A. Davis, L.C.D. Anderson, B.A. Rea, and J.A.
    Coston. Effect of adsorbed metal ions on the transport
    of Zn- and Ni-EDTA complexes in a sand and gravel
    aquifer. Geochimica et Cosmochimica Acta 66: 3017-
    3036 (2002).
Kjoller, C., D. Postma, and F. Larsen. Groundwater acidifi-
    cation and the mobilization of trace metals in a sandy
    aquifer. Environmental Science and Technology 38:
    2829-2835 (2004).
Means, J.L., D.A. Crerar, and J.O. Duguid. Migration of radio-
    active wastes: radionuclide mobilization by complexing
    agents. Science 200: 1477-1481 (1978).
McGregor, R., S. Benner, R. Ludwig, D. Blowes, and C.
    Ptacek. Sulfate reduction permeable reactive barriers
    to treat acidity, cadmium, copper, nickel, and zinc: Two
    case studies. In Handbook of Groundwater Remediation
    Using Permeable Reactive Barriers. Academic Press:
    New York, pp. 495-522 (2002).
Larsen, F. and D. Postman. Nickel mobilization in a ground-
    water well field: release by pyrite oxidation and desorp-
    tion from manganese oxides.  Environmental Science
    and Technology 31: 2589-2595 (1997).
                                                    29

-------
Lindsay, W.L.  Chemical Equilibria in Soils. New York, John
    Wiley and Sons, pp. 211 -220 (1979).
Lothenbach, B., G. Furrer, and R. Schulin. Immobilization
    of heavy metals by polynuclear aluminum and mont-
    morillonite compounds. Environmental Science and
    Technology 31: 1452-1462 (1997).
Ludwig, R., R.G. McGregor, D.W. Blowes, S.G. Benner, and
    K. Mountjoy. A permeable reactive barrier for treatment
    of heavy metals. Ground Water 40: 59-66 (2002).
Manceau, G. and G. Galas. Nickel-bearing clay minerals: II.
    Intracrystalline distribution of nickel: an x-ray absorption
    study. Clay Minerals 21: 341-360 (1986).
Manceau, G., G. Galas, and A. Decarreau. Nickel-bearing
    clay minerals: I. Optical spectroscopic study of nickel
    crystal chemistry. Clay Minerals 20: 367-387 (1985).
Manceau, A., N.Tamura, M.A. Marcus, A.A. MacDowell. R.S.
    Celestre, R.E. Sublett, G. Sposito, and H.A. Padmore.
    Deciphering Ni sequestration in soil ferromanganese
    nodules by combining X-ray fluorescence, absorption,
    and diffraction at micrometer  scales  of resolution.
    American Mineralogist 87: 1494-1499 (2002).
Manceau, A., M. Lanson, and N. Geoffrey. Natural specia-
    tion of Ni, Zn, Ba, and As in ferromanganese coatings
    on quartz using X-ray  fluorescence, absorption, and
    diffraction.  Geochimica et  Cosmochimica Acta 71:
    95-128(2007).
Mattigod, S.V., D. Rai, A.R. Felmy, and L. Rao. Solubility
    and solubility product of crystalline Ni(OH)2. Journal of
    Solution Chemistry 26: 391-403 (1997).
Merlen, E.,  P. Gueroult, J.-B. d'Espinose de la Caillerie, B.
    Rebours, C. Bobin, and O. Clause. Hydrotalcite forma-
    tion at the alumina/water interface during impregnation
    with Ni (II) aqueous solutions at neutral pH. Applied
    Clay Science 10: 45-56 (1995).
Nachtegaal, M. and D.L. Sparks. Nickel sequestration in a
    kaolinite-humic acid complex. Environmental Science
    and Technology 37: 529-534 (2003).
Nowack, B. and L.J. Sigg. Adsorption of EDTA and metal-
    EDTA complexes onto goethite. Journal of Colloid and
    Interface Science 177:  106-121  (1996).
Nowack, B., H. Xue, and L. Sigg. Influence of natural and
    anthropogenic ligands  on metal transport during inf-
    iltration of river water to groundwater. Environmental
    Science and Technology 31: 866-872 (1997).
Pye, K. Forensic examination of rocks, sediments, soils
    and dusts using scanning electron microscopy and X-
    ray chemical microanalysis. In Forensic Geoscience:
    Principles, Techniques and Applications. K. Pye and
    D.J. Croft (Eds.), Geological Society Special Publica-
    tions, Volume 232, pp.103-122 (2004).
Oakley,  D. and N.E. Korte. Nickel and chromium in  ground
    water samples as influenced by well construction and
    sampling methods. Ground Water Monitoring and Re-
    mediation 16: 93-99 (1996).
Peltier, E., R. Allada, A. Navrotsky, and D.L. Sparks. Nickel
    solubility and precipitation in soils: a thermodynamic
    study. Clays and Clay Minerals 54: 153-164 (2006).
Peltier, E., A.L. Dahl, and J.-F Gaillard. Metal speciation
    in anoxic sediments: When sulfides can be construed
    as oxides. Environmental Science and Technology 39:
    311-316(2005).
Richter, R.O. and T.L. Theis. Nickel speciation in a soil/wa-
    ter system. In Nickel in the Environment. J. O. Nriagu
    (Ed.), John Wiley & Sons, New York, NY, pp. 189-202
    (1980).
Ryan, PC., A.J. Wall, S. Hillier, and  L. Clark. Insights into
    sequential  chemical extraction procedures from quan-
    titative XRD: a study of trace  metal partitioning  in  sedi-
    ments related to frog malformities. Chemical Geology
    184:337-357(2002).
Rudd, T., J.A.  Campbell,  and J.N. Lester. Use of model
    compounds to elucidate metal forms in sewage sludge.
    Environmental Pollution 50: 225-242 (1988).
Scheckel, K.G., C.A. Impellitteri, J.A. Ryan, andT. McEvoy.
    Assessment of a sequential extraction procedure for
    perturbed lead-contaminated samples with and without
    phosphorus amendments. Environmental Science and
    Technology 37: 1892-1898 (2003).
Scheckel, K.G., A.C. Scheinost, R.G. Ford, and D.L. Sparks.
    Stability of  layered  Ni hydroxide surface precipitates - A
    dissolution  kinetics study. Geochimica et Cosmochimica
    Acta 64: 2727-2735 (2000).
Scheckel, K.G. and D.L. Sparks.  Kinetics of the formation
    and dissolution of Ni  precipitates in a gibbsite/amor-
    phous silica mixture. Journal of Colloid and Interface
    Science 229: 222-229 (2000).
Scheckel, K.G. and  D.L.  Sparks. Temperature effects on
    nickel sorption kinetics at the mineral-water interface.
    So/7 Science Society of America Journal 65: 719-728
    (2001).
Scheidegger, A.M.,  D.G. Strawn, G.M. Lamble, and D.L.
    Sparks. The kinetics of mixed Ni-AI hydroxide forma-
    tion on clays and aluminum  oxides: A time-resolved
    XAFS study.  Geochimica  et Cosmochimica Acta 62:
    2233-2245(1998).
Scheinost, A.C., R.G.  Ford, and D.L. Sparks. The role of
    Al in the formation of secondary Ni precipitates on
    pyrophyllite, gibbsite, talc, and amorphous silica: A
    DRS study. Geochimica et  Cosmochimica Acta 63:
    3193-3203(1999).
Scheinost,  A.C. and D.L. Sparks. Formation  of layered
    single and  double  metal hydroxide precipitates at the
    mineral/ water  interface: A multiple-scattering XAFS
    analysis. Journal of Colloid and Interface Science 223:
    167-178(2000).
Schultz, M.F, M.M. Benjamin, and J.F Ferguson. Adsorption
    and desorption of metals on ferrihydrite: reversibility
    of the reaction and sorption properties of the regener-
    ated solid.  Environmental Science and Technology 21:
    863-869 (1987).
                                                     30

-------
Schulze,  D.G. An introduction to soil mineralogy. In So/7
    Mineralogy with Environmental Applications. J.B. Dixon
    and D.G. Schulze (Eds.), Madison, Wisconsin, Soil Sci-
    ence Society of America, Inc., pp. 1-35 (2002).
Seaman, J.C., J.S. Arey, and P.M. Bertsch. Immobilization
    of nickel and other metals in contaminated sediments
    by hydroxyapatite addition. Journal of Environmental
    Quality 30: 460-469 (2001).
Shannon, R.D. and J.R. White. The selectivity of a sequen-
    tial extraction procedure for the determination of iron
    oxyhydroxides and  iron sulfides  in lake sediments.
    Biogeochemistry 14: 193-208 (1991).
Snodgrass, W.J. Distribution and behavior of nickel in the
    aquatic environment. In Nickel in the Environment.
    J.O. Nriagu  (Ed.), New York, John Wiley & Sons, pp.
    203-274(1980).
Stoeppler, M. Analysis of nickel in biological  materials
    and natural waters.  In Nickel in the Environment. J.O.
    Nriagu (Ed.), New York, John Wiley & Sons, pp. 661-
    821 (1980).
Strathmann, T.J.  and S.C.B. Myneni. Speciation of  aque-
    ous Ni(ll)-carboxylate and Ni(ll)-fulvic acid solutions:
    Combined ATR-FTIR and XAFS analysis. Geochimica
    et Cosmochimica Acta 68: 3441-3458 (2004).
Tessier, A., RG.C. Campbell, and M. Bisson. Sequential ex-
    traction procedure for the speciation of particulate trace
    metals. Analytical Chemistry 51: 844-851 (1979).
Thoenen, T. Pitfalls in the use of solubility limits for radioac-
    tive waste disposal: The case of nickel in sulfidic ground-
    waters. Nuclear Technology 126: 75-87 (1999).
Tonkin, J.W., L.S. Balistrieri, and J. W. Murray.  Modeling
    sorption of divalent metal cations on hydrous man-
    ganese oxide using the diffuse double layer model.
    Applied Geochemistry^: 29-53 (2004).
USEPA.  Health  Assessment  Document for  Nickel and
    Nickel Compounds, Office of Health and Environmental
    Assessment, Environmental Criteria and Assessment
    Office,  Research Triangle Park, NC, EPA-600/8-83-
    012F (1986).
USEPA. Toxic Release Inventory. Washington, DC (2003).
    http://www.epa.gov/tri/.
USEPA. 2006 Edition of the Drinking Water Standards and
    Health Advisories, EPA 822-R-06-013, Office of Water,
    Washington,  DC, (2006a).
USEPA. National Recommended Water Quality Criteria,
    4304T, Office of Water, Washington, DC, (2006b).
Van den  Berg, C.M.G.  and  M. Nimmo. Determination of
    interactions of nickel with dissolved organic material in
    seawater using cathodic stripping voltammetry. Science
    of the Total Environment 60: 185-195 (1987).
Warwick, P., A. Hall, J. Zhu,  PW. Dimmock, R. Robbins, L.
    Carlsen, and P. Lassen. Effect of temperature on the
    nickel humic acid equilibrium reaction. Chemosphere
    35:2471-2477(1997).
Wilkin, R.T.  and M.S. McNeil. Laboratory  evaluation of
    zero-valent iron to treat water impacted by acid mine
    drainage. Chemosphere 53: 715-725 (2003).
Zachara, J.M.,  J.K.  Fredrickson,  S.C.  Smith,  and  PL.
    Gassman. Solubilization of Fe(lll) oxide-bound trace
    metals by a dissimilatory  Fe(lll) reducing bacterium.
    Geochimica et Cosmochimica Acta 65: 75-93 (2001).
                                                    31

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                                                Copper
                                           Richard T. Wilkin
Occurrence and Distribution

Copper is a minor element in  the earth's crust, ranking
25th in abundance and having an average concentration of
25 mg kg~1 (Wedepohl, 1995).  Crustal copper concentra-
tions vary as a function of lithologic type and proximity to
hydrothermal deposits of copper and other base metals.
Mafic and ultramafic rocks  such as basalts are usually
more enriched in copper compared to rocks that make up
continental crust such as granites and rhyolites.  Median
concentrations of copper reported in a survey of sediments
and soils range from 7 to 35 mg kg-1 (Reimann and Caritat,
1998). Ores of copper are highly diverse and range from:
native copper deposits hosted  in sulfur-poor basaltic and
andesitic rocks; copper sulfides hosted in layered mafic
intrusions, altered porphyritic rocks, and submarine massive
sulfide bodies; and, copper oxides, carbonates, and sulfates
formed in supergene deposits  through the weathering of
primary sulfide deposits (Guilbert and Park, 1986).

It is believed that the ancient civilizations of Mesopotamia
(region of modern Iraq) made use of native copper and gold,
possibly as far back as ca. 8700  BC. Indeed the exploitation
of natural deposits of copper and gold, which are soft met-
als that can be  hammered into shape without heat, marked
the transition from the Stone Age to more modern ways of
life (Diamond, 1997). Typical copper-bearing minerals in-
clude native copper, sulfides (chalcopyrite, CuFeS2; bornite,
Cu5FeS4; chalcocite, Cu2S;covellite, CuS;digenite, CugS5),
sulfosalts (tetrahedrite, Cu12Sb4S13), carbonates (malachite,
Cu2(OH)2CO3; azurite, Cu3(OH)2(CO3)2), and oxides (teno-
rite, CuO). In the US, the principal copper mining states are
Arizona, Utah,  and New Mexico.  In 2004, domestic mine
production of copper was about 1.16 million tons, or about
8% of world mine production. While the US is the world's
second-largest producer of copper, it is the world's largest
copper-consumer. The principal modern use of copper is
as an electrical conductor.  Alloys of copper (e.g., brass
and bronze) are used in jewelry, sculptures and for mint-
ing coins. Copper also has broad uses as an agricultural
poison and as an algaecide in water purification.

Plume Characteristics

Copper has five possible oxidation states (0, +1, +2, +3,
and  +4).  Under  most conditions,  copper is present in
aqueous solution as the divalent cation,  Cu2+, or as Cu(ll)
hydroxide or carbonate complexes.  However, copper is
not especially mobile in aquatic environments due to the
relatively low solubility of Cu(ll)-bearing solids  and high
affinity of copper for mineral and organic surfaces.  Cer-
tain organic compounds are able to keep copper soluble
by ligand complexation. Hence, copper does not typically
enter ground water except under conditions of low pH or
high ligand concentrations.  Where present, copper con-
tamination in soils and ground water stems  primarily from
mining activities, metal production, wood production, fertil-
izer production, and combustion of fossil fuels and wastes
(e.g., Bochehska et al., 2000; Zagury et al.,  2003).

Copper is not included on  the CERCLA Priority List of
Hazardous Substances, which  is based on  the frequency
of occurrence of specific contaminants at National Priorities
List (NPL) sites and their potential threat to human health.
An internet search showed,  however, that in 2005 copper
was listed as a potential contaminant of concern (COG) in
ground water at 287 NPL sites  in EPA Regions 1-10.

Remedial Technologies

The primary techniques for  dealing with copper-contami-
nated soils involve immobilization and/or extraction. Immo-
bilization involves binding copper or other heavy metals to
the soil matrix by solidification or stabilization. In this way,
contaminated soils become less soluble, and hazardous
compounds are prevented from entering ground water or
surface water. Extraction involves a combination of process-
es to actually remove heavy metals from soil, for example,
soil washing whereby metals are transferred into solution
via solubilization by acids, bases, or chelating agents. In
contrast to soils remediation, there are comparatively few
examples of  ground-water  remediation demonstrations
that  focus on copper.  In situ bioremediation to promote
bacterial sulfate-reduction and  consequent precipitation
of insoluble copper sulfides has been proposed (Dvorak et
al.,  1992; Steed et al., 2000; Tabak et al., 2003).  Perme-
able reactive barriers that are  designed  to  intercept  and
treat contaminated ground water could be appropriate for
dealing with copper contamination. Woinarski et al. (2003)
discuss the application of a natural zeolite (clinoptilolite) in
reactive barriers for removing copper via ion exchange (see
also Inglezakis et al., 2003 and Park et al.,  2002). Other
reactive media explored in laboratory studies for treating
copper include zerovalent iron  (Wilkin and McNeil, 2003)
and municipal compost (Waybrant et al., 1998).
                                                    33

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Regulatory Aspects

The USEPA has set the Maximum Concentration Limit Goal
(MCLG) of copper in drinking water at 1.3 mg L1 (USEPA,
2006a; http://www.epa.gov/waterscience/criteria/drinking/
dwstandards.pdf). Copper is a trace element essential for
good human health.  It is part of the prosthetic groups of
many proteins and enzymes and thus is essential to their
proper function.  Potential  health effects from  ingesting
water with high concentrations of copper include gastro-
intestinal distress, and potential damage to the liver and
kidneys.  For non-potable water sources, ambient water
quality  criteria  (AWQC) that are protective of aquatic life
may serve as alternative cleanup goals. For copper, current
statutes list both acute and chronic criteria for fresh waters
as 0.016 mg L1 and 0.011 mg L1,  respectively, for a water
hardness of  100 mg  L1  (USEPA,  2006b; http://www.epa.
gov/waterscience/criteria/nrwqc-2006.pdf).  Adjustments
to these criteria are to be applied for waters with different
hardness. An example of where this criterion may apply
is a site where contaminated ground water discharges to
surface water.

Geochemistry and Attenuation Processes

Aqueous Speciation

Copper complexes are possible in the +1, +2, +3, and +4
valence states.  Copper(lll)  and (IV) complexes are rare
and unstable in  water. Cu(l) complexes are present under
reducing conditions but in general cuprous ions are highly
insoluble in water. Cu(ll) is the main oxidation state for
soluble complexes of copper in aquatic environments. Cu(ll)
forms complexes with both  hard (e.g.,  CO32-, SO42-, OH-,
and Ch) and  soft (e.g., S2~, h) bases  (Stumm and Morgan,
1996).  The stereochemistry of Cu(ll) principally involves
distorted tetragonal (coordination number 4) or octahedral
(coordination number 6)  configurations.  Cu(ll) complexes
are subject to the Jahn-Teller effect that acts to stabilize
species with the d9 electronic configuration that are present
in tetragonal or  octahedral coordination.

In pure water,  Cu2+ is the predominant ion below pH 7.
Above this pH, the species CuOH+, Cu(OH)2, and CuO22-
become increasingly  important. Because of uncertainties
in the estimates and measurements of the thermodynamic
constants for copper complex formation, the speciation
of copper in natural  waters is  not known in detail (e.g.,
Boyle,  1979; Leckie and Davis,1979; Baes and Mesmer,
1976).  Most models,  however, predict that Cu2+ is a small
fraction of the total copper concentration in freshwater and
seawater systems and that complexed forms of copper are
dominant.

A survey of references that report  hydrolysis constants for
Cu2+ shows considerable variability in species identified and
in their formation constants  (see e.g., Baes and Mesmer,
1976; Leckie and  Davis, 1979 and references  therein).
Figure 4.1 shows the pH-dependent distribution of Cu2+ hy-
drolysis species based on the Lawrence Livermore National
Laboratory thermodynamic database (thermo.com.v8.r6+),
along with the pH-dependent solubility of tenorite (CuO) and
the metastable solid, Cu(OH)2, based on data in Hidmi and
Edwards (1999). Both CuO and Cu(OH)2 are insoluble at
neutral to alkaline pH. Below pH of 6 to 7, dissolution of
these phases would yield Cu2+ concentrations  above the
MCL of 1.3 mg L1 (1Q-47 molal). Stable complexes of Cu2+
with  SO42- (CuSO4) and CO32- (CuCO3) may  contribute
significantly to total copper at anion concentrations typi-
cally encountered in ground water.  In  environments with
high  ammonia concentrations, copper can be significantly
bound to ammonia at pH>6.

Copper may be strongly complexed by dissolved organic
matter (e.g.,  Smolyakov et al., 2004).  Low  molecular
weight, dissolved organic complexes are generally  highly
mobile and able to transport copper in aquifer materials
and soils (Han and Thompson, 2003; Christensen  et al.,
1999). Christensen et al. (1999) report that >85% of total
copper was bound to dissolved organic carbon complexes
in leachate with comparatively low dissolved organic carbon
concentrations of <40 mg C L~1.
    -2
 d
 CD
 03
 O
    -10
    -12
    -14
                                     10
                                            12     14
                           PH
Figure 4.1  Solubility of copper oxide and copper
            hydroxide as a function of pH in the system
            Cu-O-H at 25 C.

Solubility

An  Eh-pH diagram for copper  is shown in Figure 4.2.
Inspection of this diagram indicates that at the specified
conditions Cu2+ is the soluble form of copper at pH<6 and
in moderately to  highly oxidizing systems. Consequently,
upon weathering, copper is likely to be more mobile under
acidic rather than alkaline conditions (see,  e.g.,  Paulson
and Balistrieri, 1999). Copper hydroxycarbonate (mala-
chite) has a narrow stability field at near-neutral pH and at
moderately to highly oxidizing conditions. With increasing
inorganic carbon concentrations, the malachite stability field
                                                    34

-------
would expand. At neutral to alkaline pH (>7) copper oxides
are stable.  With progressively more reducing conditions,
cuprous oxide and elemental copper develop broad stabil-
ity fields. Finally, in highly reducing and sulfidic environ-
ments, copper su If ides (chalcocite and covellite) are stable
over a wide pH range. Aquifer materials usually contain
some organic matter as well as sulfate from ground water.
Microbial degradation of organic matter can be coupled to
sulfate reduction with the production of  hydrogen sulfide.
Because of the extreme insolubility of copper sulfides, no
complexing ligand can compete with  hydrogen  sulfide or
metal sulfide surfaces for copper (Rose, 1989).  In addi-
tion, there  are  several fairly common copper-iron-sulfur
minerals such as chalcopyrite (CuFeS2), bornite (Cu5FeS4),
and cubanite (CuFe2S3).  Note that in solutions with high
chloride concentrations (>1 M), the field  of soluble copper
in Figure 4.2 expands  substantially due to the very strong
nature of copper chloride complexes.
   0.5
 LU
  -0.5
       25C
      0      2     4      6      8     10     12     14
                           PH


Figure 4.2  Eh-pH diagram for copper at 25 C (total
            inorganic carbon = 10~2 molal; total sulfur =
            10~3 molal; total copper = 10~5 molal).

Cavallaro and McBride (1980) found that in alkaline soils
copper was present as Cu(OH)2 and  with progressive
aging, copper solubility decreased consistent with the for-
mation of Cu2(OH)2CO3.  In general, they concluded that
Cu2+ solubility in soil is highly correlated to pH. Dudley et
al. (1991) proposed the formation of CuO in soil reacted
with an extract of acid mine waste.  In both the studies of
Cavallaro and  McBride (1980)  and Dudley et  al. (1991),
proposed copper phase associations are based on a com-
parison of observed pH-dependent copper concentrations
with solubility estimates  based on thermodynamic data.
Indeed, Leckie and Davis (1979) suggest that in most soil
environments malachite and tenorite are the most impor-
tant copper-bearing phases, with  Cu(OH)2 present as a
metastable precursor to malachite and tenorite.

The stability relationships between copper hydroxycarbon-
ates and oxides are shown in Figure 4.3 in terms of pH
and CO2 fugacity. Note that at high pH and CO2fugacity,
aqueous copper carbonate complexes predominate over
hydroxyl complexes. The diagram  illustrates that over the
pH and f CO2 conditions in most ground water systems,
tenorite and  malachite are the expected stable copper
minerals.

Adsorption

McBride  and Bouldin (1984)  examined  the solid-phase
properties of  copper in copper-contaminated soil. They
concluded that long-term reaction of copper with calcare-
ous soil failed to convert copper into a form unavailable to
plants.  Chemical extraction tests  suggested that copper
was  mainly present in a non-exchangeable form easily
dissolved by organic chelating compounds. An analysis
of pH-dependent solubility  data revealed  that copper was
present as an inorganic form in the soil, possibly tightly
adsorbed on surfaces as hydroxyl or hydroxycarbonate spe-
cies. Greater than 99.5% of the copper in the soil solution
was complexed, probably with soluble organic compounds
(McBride and Bouldin, 1984). In contrast, Cavallaro and
McBride  (1978) found that  low pH soils are less effective
in retaining Cu2+ compared to neutral soils and calcareous
soils. They concluded that this behavior was in part related
to increased competition at low pH for organic functional
groups by aluminum and/or  protons thus reducing the ability
of Cu2+ to be adsorbed onto solid organic matter.
   -10
Figure 4.3  Solubility and speciation of copper as a
            function of pH and log fugacity of CO2(gas)
            at 25 C. Solid lines separate stable phases
            and dotted lines separate aqueous species
            (total copper = 1Q-5 molal). Diagram drawn
            using thermodynamic data from MINTEQA2.
                                                    35

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Based on the solubility and sorption behavior of copper, it is
expected that over a wide range of geochemical conditions
copper will be effectively stable in the solid phase of soils
and sediment materials. Copper has a strong affinity for
the surfaces of iron oxides and hydroxides (e.g., Benjamin
and Leckie,  1981; Robertson and Leckie, 1998; Martmez
and McBride, 1998), clays (e.g., Pickering, 1980; Farqhar
et al., 1997;  Morton et al., 2001), sulfides (e.g., Pattrick et
al., 1997; Parkman et al., 1999), and organic matter (e.g.,
Sauve et al., 1997; Schilling and Cooper, 2004). As well
as being less soluble, Cu2+ is more strongly adsorbed to
mineral substrates than Zn2+, Ni2+, and Cd2+.
Benjamin and Leckie (1981) examined the pH-dependent
sorption of copper, zinc, and lead onto hydrous ferric oxide.
For these metals and for a range of precipitate loadings,
the adsorption edge position,  the pH  at which  half the
metal was sorbed and half the metal remained in solution,
increased in  the order Pb6,
copper was essentially completely removed from solution.
Khaodhiar et al. (2000) observed nearly identical pH-depen-
dent behavior for copper adsorption onto iron oxide coated
sand grains. Martmez  and Motto (2000) determined the
pH at which metal  amended soils began to release cop-
per via an acid titration method. Interestingly, they found
that copper was released at about pH 5.5  0.2 which is in
good agreement with the adsorption studies, and further
reinforces the notion of reversible sorption processes and
potential copper mobility at low pH.

Redox Chemistry
Equilibrium between cupric and cuprous ions can be rep-
resented by  the equation:
               Cu2+ + e- = Cu+
(log K= 2.72)
In natural systems the stable solid in very reducing condi-
tions is expected to be cuprous sulfide (Cu2S, chalcocite,
see Figure 4.2). As  the Eh increases there is a narrow
window in which cupric sulfide (CuS, covellite) becomes
important. Further increases in Eh can lead to the forma-
tion of elemental copper.  So in general the solubility and
speciation of copper are  determined by redox equilibria
of sulfur and copper and the strength of available ligands.
Experimental studies of Cu(l) complexation by chloride and
bisulfide are presented in Xiao et al. (1998), Thompson and
Helz (1994), Mosselmans et al. (1999), and Mountain and
Seward (1999, 2003). Luther et al. (2002)  show that the
reduction of Cu(ll) to Cu(l) occurs in sulfidic solutions prior
to the precipitation of copper sulfides.

Colloidal Transport
Recent studies are consistent in demonstrating that copper
in ground water is frequently associated with colloids that
appear to be organic in nature (Sanudo-Wilhelmy et al.,
2002; Jensen et al., 1999; Freedman et al., 1996; Pauwels
et al., 2002). The association between metals and ground
water colloids is evident both in uncontaminated ground-
water (e.g.,  Sanudo-Wilhelmy et al., 2002) and  in  con-
taminated landfill leachates and in ground water impacted
by mining districts (e.g., Jensen et al., 1999; Pauwels et
al., 2002). Jensen et al. (1999) found that 86-95% of total
copper in landfill leachate was associated with small-size
colloidal matter and organic molecules. They concluded
that most metals, including copper, present in the colloidal
forms would have been sampled in the dissolved fraction
if the commonly employed filter size of 0.45 urn had been
used, since only negligible amounts of metal were found
with colloids >0.40 urn.  Pauwels et al. (2002) found that
the  mobility of copper  in ground water impacted by the
oxidative dissolution of  massive  sulfide deposits  in the
Iberian Pyrite Belt (Spain) was especially enhanced due
to complexation with organic matter and/or adsorption onto
colloids. For example, measured concentrations of copper
were 10s to 109 times greater than concentrations modeled
assuming equilibrium with respect to sulfide minerals (e.g.,
chalcopyrite).

Site Characterization

Copper mobility  in ground water and the risk of copper
exposure to plants, animals, and/or humans is governed by
the total concentration of copper, the distribution of copper
species in water, and the nature of copper partitioning in the
solid phase. The development of site-conceptual models
for predicting the long-term fate of copper at a contami-
nated site will require information on the distribution and
concentration of copper in the aqueous phase and the solid
phase. Table 4.1 indicates possible natural attenuation and
mobilization pathways for copper.

Quantitative measurement of  copper  concentrations in
aqueous solutions is typically carried out using inductively
coupled plasma optical emission spectroscopy (ICP-OES),
inductively coupled plasma mass spectroscopy (ICP-MS),
or atomic absorption spectroscopy (AAS).   Some  of the
unique features of determining copper concentrations in
natural waters are discussed in  Boyle (1980)  and Sanudo-
Wilhelmy et al. (2002).  Input data to geochemical  codes
(e.g.,  MINTEQA2,   PHREEQC, EQ3/6) for determining
aqueous speciation  also require, at a minimum, the con-
centrations of major anions,  major cations,  total organic
carbon, temperature, and pH.  The total concentration of
copper in soils, sediments, and aquifer materials may be
determined by X-ray fluorescence (XRF) spectroscopy, or
by chemical analysis after digestion in mineral acids.

While  regulatory requirements stipulate  that unfiltered
ground-water samples  be analyzed to support  regula-
tory decisions at a  contaminated site, it may be neces-
sary to also collect filtered samples to  help  interpret that
process(es) controlling  contaminant mobility. The use of
0.45 urn pore size filter paper is common as an  arbitrary
cutoff point to differentiate between dissolved and particu-
late phases in water samples. However, caution is recom-
mended when using this approach, particularly for Fe and
Al and other elements that may be associated with Fe or
Al particles (including Cu) that could pass through 0.45 um
filter papers. The use of filter papers with pore sizes less
than 0.1 um will generally provide a better assessment of
the dissolved vs. particulate load of a ground water or a
surface water sample.
                                                     36

-------
Mickey and Kittrick (1984) examined the chemical partition-
ing of copper in soils and sediments containing high levels
of heavy metals using the selective extraction approach
developed by Tessier  et al. (1979).  In this study, copper
was  assigned  to five  operationally defined geochemical
fractions: exchangeable, bound to carbonates, bound to
Fe- and Mn-oxides, bound to organic matter, and residual.
This  study concluded that copper was the metal most
significantly associated with organic matter.  Compared to
other heavy metals considered in this study (nickel, and
zinc), copper displayed a  low potential for mobility and
metal bioavailability (Mickey and Kittrick, 1984).

Long-Term Stability and Capacity

The stability of attenuated copper will largely depend on the
fluctuation of site-specific geochemical conditions through
time.  For example, if copper attenuation follows a copper
hydroxycarbonate precipitation  pathway, then  long-term
stability of attenuated  copper will depend, in part,  on the
persistence of pH conditions. If pH conditions were to shift
significantly to more acidic values, copper might be expect-
ed to release from the solid phase. It is therefore important
to understand the attenuation mechanism(s) so that geo-
chemical triggers for remobilization can be anticipated and
incorporated into evaluations of long-term monitoring data.
For any proposed and identified attenuation mechanism,
there will exist possible scenarios whereby remobilization
can occur (i.e., changes in pH or Eh). It will be essential to
explore the likelihood of such changes in site geochemistry
and the sensitivity of the attenuation pathway to changes
in the prevailing geochemical conditions.

Quantifying the attenuation capacity (as defined in Volume
1) will also necessitate an  understanding of the specific
attenuation pathway(s). Attenuation capacity, for example,
could be related to the extent that pH is  buffered,  the
availability of sorptive sites in aquifer materials, or to the
supply of electron  donors  needed to sustain microbially
                         mediated redox conditions. For any proposed attenuation
                         mechanism, there will be assumptions built into capacity
                         estimations, so that uncertainty analysis is recommended
                         to support capacity calculations.

                         Tiered Analysis

                         Determination  of the viability of copper  remediation in
                         ground water via monitored natural attenuation will depend
                         upon proper assessment of  contaminant loading to the
                         aquifer and prevailing geochemistry and mineralogy within
                         the contaminant plume and the down gradient zone prior
                         to the point(s) of compliance.  MNA may not be appropri-
                         ate as a site remedy for copper contamination in acidic
                         pH, highly oxidizing, and/or DOC-rich environments. The
                         goal of site assessment is to demonstrate the process(es)
                         controlling copper sequestration onto aquifer solids and
                         the long-term stability of solid phase copper as a function
                         of existing and anticipated ground-water chemistry. The
                         following tiered analysis structure for site characterization
                         provides an approach to evaluate candidate sites and de-
                         fine the potential limitations of MNA as part of a remedy
                         for ground-water cleanup.

                         Tier I - Site characterization under Tier I will involve dem-
                         onstration that the plume is  static or shrinking, has not
                         reached compliance boundaries,  and does  not  impact
                         existing water supplies. Once this is established through
                         ground-water  characterization, evidence  is  collected to
                         demonstrate Cu  partitioning  to aquifer solids within the
                         plume. If natural attenuation processes are active through-
                         out the plume, then there should be an observed increase
                         in solid phase concentrations within regions of the plume
                         with higher aqueous concentrations, e.g., near the source
                         term.  This field partitioning data may be supplemented
                         by geochemical  modeling that incorporates  measured
                         water chemistry (e.g.,  pH, Eh, and major ion chemistry)
                         throughout the plume to assess the potential for solubility
                         control by copper hydroxide, sulfate, carbonate, phosphate,
Table 4.1  Natural attenuation and mobilization pathways for copper.
    Attenuation Processes
     Mobilization Processes
        Characterization Approach
 Precipitation of insoluble hy-
 droxycarbonates, carbonates,
 sulfides, and phosphates.  In
 general, pH>6 will drive pre-
 cipitation reactions resulting in
 Cu concentrations below the
 MCL.
Dissolution of carbonates, hydroxy-
carbonates, and oxides at low pH;
oxidative dissolution of sulfides
at low pH and high Eh; complex-
ation/stabilization  in the presence
of DOC.
Evaluation of copper speciation in the aque-
ous phase. Determination of total Cu in the
solid matrix.  Evaluation of solid phase parti-
tioning using  sequential extraction methodolo-
gies.  Evaluation of long-term sorption capac-
ity/stability.
 Sorption to iron hydroxides,
 organic matter, carbonates,
 and sulfides.
Desorption at low pH; complex-
ation/stabilization in the presence
of DOC.  Reductive dissolution of
iron hydroxides.
Evaluation of copper speciation in the aque-
ous phase. Determination of total Cu in the
solid matrix.  Evaluation of solid phase parti-
tioning using  sequential extraction methodolo-
gies.  Batch and column testing to determine
Cu uptake capacity of site-specific aquifer ma-
terials with variable geochemical conditions.
                                                     37

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or sulfide. This provides justification for proceeding to Tier
II characterization efforts.

Tier II - Under Tier II, the apparent rate and mechanism(s)
of attenuation are determined.  Estimates of a site attenu-
ation rate(s)  can  be assessed via a well transect  along
the ground-water  flow path.  In addition, time-series data
may be collected  at one or more monitoring points  within
the plume. This information  will allow assessment  of the
relative timescales for contaminant immobilization and fluid
transport and determination of whether remediation objec-
tives can be met within the required regulatory time frame.
In addition, the mechanism(s) for attenuation need to be
identified under this stage of site characterization.  This
effort may require  determination of the chemical speciation
of aqueous and solid phase Cu, which may be approached
according to the following scheme:
 1)
 2)
     Determination of solution speciation via direct ana-
     lytical measurements (e.g., Martinez et al., 2001;
     Sanudo-Wilhelmy et al., 2002) to aid differentiation
     of uncomplexed (i.e., Cu2+) and complexed (e.g.,
     CuO, Cu-organic ligand complexes) forms of mo-
     bile Cu in combination with speciation calculations
     based on characterized ground-water chemistry;
     Calculation of the saturation state of ground water
     relative to measured aqueous  chemistry compli-
     mented  by the possible isolation of  discrete Cu
     mineral phases via density separations (or other
     schemes) in regions of the aquifer with highest
     solid phase concentrations;
 3)
     Determination of aquifer mineralogy to determine
     the relative abundance of components with docu-
     mented capacity for Cu sorption (e.g., Amonette,
     2002; Burton et al., 2005); and
 4)   Determination  of  Cu-sediment associations via
     chemical  extractions designed  to target specific
     components within the aquifer sediment (e.g., Lee
     etal., 2005).
This compilation of  information will facilitate identification
of the reaction(s)  leading to Cu immobilization within the
plume.

Tier III -  Once the partitioning mechanism(s) have been
identified for the site, the subsequent characterization effort
under Tier III will  involve determination of the stability of
immobilized Cu and the capacity of the aquifer to sustain
continued uptake.  It is recommended that the stability of
immobilized Cu be tested based on the anticipated evolution
of ground-water chemistry concurrent with plume shrinkage.
For example, changes in ground-water pH can exert a sig-
nificant influence on Cu adsorption or precipitate solubility.
Therefore, it is recommended that sediment leach tests be
conducted to characterize the magnitude of Cu mobilization
as a function of pH for a ground-water chemistry representa-
tive of site conditions. It is recommended that the capacity
for Cu uptake onto aquifer solids be determined relative to
the specific mechanism(s) identified in Tier II. For example,
if site characterization under Tier II indicated that precipita-
tion of Cu sulfide due to microbial degradation of organic
compounds coupled with sulfate reduction occurs within the
aquifer, then it is recommended that the mass distribution
of organic carbon and sulfate to support this reaction within
the aquifer be determined. This site-specific capacity can
then be compared to  Cu mass loading within the plume
in order to assess the longevity of the natural attenuation
process. If site-specific tests  demonstrate the stability of
immobilized Cu and sufficient capacity within the aquifer to
sustain Cu attenuation, then the site characterization effort
can progress to Tier IV. For cases where contaminant stabil-
ity is sufficient but aquifer capacity is insufficient for capture
of the entire plume, then a determination of the benefits of
contaminant source reduction may be necessary.

Tier IV - Finally,  under Tier IV a monitoring plan is estab-
lished along with contingency plans in the event of MNA
failure.  It is  recommended that the monitoring  plan be
designed to establish both continued plume stability and to
identify changes in ground-water chemistry that may lead
to re-mobilization of attenuated Cu. The specific chemical
parameters to be monitored  will include those identified
under Tier III  that may halt Cu partitioning to aquifer sedi-
ments and/or result in solubilization of either discrete Cu
precipitates or aquifer minerals that sequester Cu  from
ground water. For  example,  solution phase  parameters
that could alter  either Cu precipitation or adsorption in-
clude increases  in soluble organic carbon or chloride in
combination with changes in ground-water pH. In contrast,
the concentration of dissolved iron or sulfate may indicate
the dissolution of an important sorptive phase within the
aquifer (e.g.,  reductive dissolution of iron oxides or oxida-
tive dissolution of sulfides).  Changes in these parameters
may occur prior to observed changes in solution Cu  and,
thus, serve as monitoring triggers for potential MNA failure.
In this instance,  a contingency plan can be implemented
that incorporates interventive  strategies to arrest possible
plume expansion beyond compliance boundaries. Possible
strategies to prevent plume expansion include pump and
treat operations,  installation of reactive barriers to enhance
uptake capacity  perpendicular to the direction of plume
advance, or enhancement of natural attenuation processes
within the aquifer through the injection of soluble reactive
components.

References

Amonette, J.E.  Methods for determination  of mineral-
    ogy and environmental availability. In  So/7 Mineralogy
    with Environmental Applications.  J.B. Dixon and  D.G.
    Schulze (Eds.),  Soil Science Society of America Book
    Series. Soil Science Society of America, Inc., Madison,
    Wl, pp. 153-197(2002).
Baes, C.F, and R.E. Mesmer. The Hydrolysis of Cations.
    Wiley Interscience, New York, NY, 489 pp. (1976).
Benjamin, M.M.,  and J.O. Leckie. Multiple-site adsorption
    of Cd, Cu, Zn,  and Pb on amorphous iron hydroxide.
    Journal of Colloid and Interface Science 79: 209-221
    (1981).
                                                     38

-------
Bochehska, T., J. Fiszer, and M.  Kalisz.   Prediction  of
    groundwater  flow  into copper mines  of the Lubin
    Gtogow Copper District.  Environmental Geology 39:
    587-594 (2000).
Boyle, E.A.  Copper in natural waters. In Copper in the
    Environment. J.O. Nriagu (Ed.), Wiley, New York, NY
    pp. 77-88(1979).
Burton, E.D.,  I.R. Phillips, and D.W. Hawker. Geochemi-
    cal partitioning of copper, lead, and zinc in benthic,
    estuarine sediment profiles. Journal of Environmental
    Quality 34: 263-273  (2005).
Cavallaro, N. and M.B. McBride.  Activities of Cu2+ in soil
    solutions  as affected by pH.  So/7 Science Society of
    America Journal 44: 729-732 (1980).
Cavallaro, N.  and M.B. McBride.   Copper and cadmium
    adsorption characteristics of selected acid and calcare-
    ous soils. So/7 Science Society of America Journal 42:
    550-556 (1978).
Christensen, J.B., J.J. Botma, andT.H. Christensen. Com-
    plexation of Cu and Pb by DOC in polluted groundwater:
    A comparison of experimental data and predictions by
    computer speciation models (WHAM and MINTEQA2).
    Water Research 33: 3231-3238 (1999).
Diamond, J. Guns, Germs, and Steel:  The Fate of Human
    Societies. W.W.Norton,  New York, NY (1997).
Dudley, L.M., J.E.  McLean,  T.H. Furst, and J.J. Jurinak.
    Sorption of cadmium and copper from  an acid mine
    waste extract by two calcareous soils: Column studies.
    So/7 Science 151: 121-135 (1991).
Dvorak, D.H.,  R.S. Hedin,  H.J. Edenborn, and PE. Mclntire.
    Treatment of metal-contaminated water using bacterial
    sulfate reduction: Results from pilot scale reactors. Bio-
    technology and Bioengineering 40: 609-616 (1992).
Farquhar, M.L., D.J. Vaughan, C.R. Hughes, J.M. Charnock,
    and K.E.R. England. Experimental studies of the inter-
    action of aqueous metal cations with mineral substrates:
    Lead, cadmium, and copper with perthitic  feldspar,
    muscovite, and biotite. Geochimica et Cosmochimica
    Acta  61: 3051 -3064 (1997).
Freedman, Y.E., D.  Ronen, and G.L. Long.  Determination
    of Cu and Cd content of groundwater colloids by sam-
    pling  graphite furnace atomic absorption spectrometry.
    Environmental Science and Technology 30:2270-2277
    (1996).
Guilbert,  J.M. and C.F. Park.  The Geology of Ore Deposits.
    W. H. Freeman and Company,  New York, NY 985 pp.
    (1986).
Han, N. and M.L. Thompson. Impact of dissolved organic
    matter on copper mobility in aquifer material. Journal
    of Environmental Quality 32:  1829-1836 (2003).
Hickey, M.G.  and  J.A. Kittrick.  Chemical partitioning  of
    cadmium, copper,  nickel and zinc in  soils and sedi-
    ments containing high levels of  heavy metals. Journal
    of Environmental Quality 13: 372-376  (1984).
Hidmi, L. and M. Edwards.  Role of temperature and pH in
    Cu(OH)2 solubility. Environmental Science and Technol-
    ogy 33:2607-2610 (1999).
Inglezakis, V.J., M.D. Loizidou, and H.P Grigoropoulou. Ion
    exchange of Pb2+, Fe3+, and Cr3+ on natural clinoptilolite:
    Selectivity determination and influence of  acidity on
    metal uptake. Journal of Colloid and Interface Science
    261: 49-54 (2003).
Jensen, D.L., A. Ledin, and T.H. Christensen. Speciation of
    heavy metals in landfill-leachate polluted groundwater.
    Water Research 33: 2642-2650 (1999).
Khaodhiar, S., M.F Azizian, K. Osathaphan, and P.O. Nel-
    son.  Copper, chromium, and arsenic adsorption and
    equilibrium  modeling  in  an  iron-oxide-coated  sand,
    background electrolyte system.  Water, Air, and Soil
    Pollution 119: 105-120  (2000).
Leckie, J.O. and J.A. Davis.  Aqueous environmental chem-
    istry of copper.  In  Copper in the Environment. J.O.
    Nriagu, (Ed.), Wiley, New York, NY pp. 89-121 (1979).
Lee, P.-K., Y.-H. Yu, S.-T. Yun, and B. Mayer.  Metal con-
    tamination and solid phase  partitioning of metals in
    urban roadside sediments. ChemosphereGO: 672-689
    (2005).
Luther, G.W., S.M.Theberge, T.F Rozan,  D. Rickard, C.C.
    Rowlands, and A. Oldroyd.  Aqueous copper sulfide
    clusters  as  intermediates during copper sulfide for-
    mation.  Environmental Science and Technology 36:
    394-402 (2002).
Martinez, C.E. and M.B. McBride. Solubility of Cd2+, Cu2+,
    Pb2+, and Zn2+ in aged coprecipitates with amorphous
    iron hydroxides.  Environmental Science and Technol-
    ogy 32: 743-7'48 (1998).
Martinez, C.E. and H.L. Motto. Solubility of lead, zinc, and
    copper added to mineral soils. Environmental Pollution
    107: 153-158(2000).
Martinez, C.E., A. Jacobson, and  M.B. McBride.  Thermally
    induced  changes in metal solubility of contaminated
    soils is linked to mineral recrystallization and organic
    matter transformations.  Environmental Science and
    Technology 35: 908-916 (2001).
McBride, M.B. and D.R. Bouldin.  Long-term reactions of
    copper(ll) in a contaminated soil. So/7 Science Society
    of America Journal 48: 56-59 (1984).
Morton, J.D., J.D. Semrau, and K.F Hayes. An X-ray absorp-
    tion spectroscopy study of the structure and reversibility
    of copper adsorbed to montmorillonite clay. Geochimica
    et Cosmochimica Acta 65: 2709-2722 (2001).
Mosselmans, J.F.W., R.A.D. Pattrick, J.M. Charnock, and
    V.A. Sole. EXAFS of copper in  hydrosulfide solutions
    at very low concentrations: Implications for  speciation
    of copper in natural waters.  Mineralogical Magazine
    63:769-772 (1999).
Mountain, B.W. and T.M.  Seward.  The  hydrosulphide/
    sulphide complexes of  copper(l): Experimental deter-
                                                    39

-------
    mination of stoichiometry and stability at 22  C and
    reassessment of high temperature data.  Geochimica
    et Cosmochimica Acta 63:11 -29 (1999).
Mountain, B.W. andT.M. Seward. Hydrosulphide/sulphide
    complexes of copper(l): Experimental confirmation of
    the stoichiometry and stability of Cu(HS)2" to elevated
    temperatures. Geochimica et Cosmochimica Acta 67:
    3005-3014(2003).
Park, J.-B., S.-H. Lee, J.-W. Lee, and C.-Y. Lee. Lab scale
    experiments for permeable reactive barriers against
    contaminated groundwater with ammonium and heavy
    metals using clinoptilolite.  Journal of Hazardous Ma-
    terials B95: 65-79 (2002).
Parkman, R.H., J. M. Charnock, N.D. Bryan, F.R. Livens,
    and D.J.Vaughan.  Reactions of copper and cadmium
    ions in aqueous solution with goethite, lepidocrocite,
    mackinawite, and pyrite.  American Mineralogist 84:
    407-419 (1999).
Pattrick, R.A.D., J.F.W. Mosselmans, J.M. Charnock,  K.E.R.
    England, G.R. Helz,  C.D. Garner,  and D.J. Vaughan.
    The structure of amorphous copper sulfide precipitates:
    An X-ray absorption study. Geochimica et Cosmochi-
    mica Acta 61: 2023-2036 (1997).
Paulson, A.J. and L. Balistrieri.  Modeling removal  of Cd,
    Cu, Pb, and Zn in acidic groundwater during neutral-
    ization by ambient surface waters  and groundwaters.
    Environmental Science and Technology 33:3850-3856
    (1999).
Pauwels, H., M.Tercier-Waeber, M. Arenas, R. Castroviejo,
    Y. Deschamps, A. Lassin, F. Graziottin, and F. Elorza.
    Chemical characteristics of groundwater around two
    massive sulphide deposits in an area of previous min-
    ing contamination, Iberian Pyrite Belt, Spain. Journal
    of Geochemical Exploration 75: 17-41 (2002).
Pickering, W.F Copper retention by soil/sediment compo-
    nents. In Copper in the Environment. J.O. Nriagu (Ed.),
    Wiley, New York, NY  pp. 217-253 (1980).
Reimann, C. and P. Caritat.  Chemical Elements in the En-
    vironment.  Springer-Verlag, Berlin, 398 pp. (1998).
Robertson, A.P. and J.O.  Leckie. Acid/base, copper bind-
    ing, and Cu2+/H+ exchange properties of  goethite, an
    experimental and modeling study.  Environmental Sci-
    ence and Technology 32: 2519-2530 (1998).
Rose, A.W.  Mobility of  copper and other heavy metals
    in sedimentary Environments.  In Sediment-hosted
    Stratiform Copper Deposits. R.W. Boyle,  A.C. Brown,
    C.W. Jefferson, E.G. Jowett, and R.V.  Kirkham (Eds.),
    Geological Association of Canada, Special Paper 36,
    pp. 97-110(1989).
Sanudo-Wilhelmy, S.A., F.K. Rossi,  H.  Bokuniewicz, and
    R.J. Paulsen.  Trace  metal levels  in uncontaminated
    groundwater of a coastal  watershed: Importance of
    colloidal forms. Environmental Science and Technology
    36: 1435-1441 (2002).
Sauve, S.,  M.B. McBride, W.A.  Norvell, and W.H. Hen-
    dershot.  Copper solubility and speciation of in situ
    contaminated soils: Effects of copper level,  pH and
    organic matter.   Water, Air, and  Soil Pollution 100:
    133-149 (1997).
Schilling, M. and W.T Cooper.  Identification of copper
    binding sites in soil organic matter through chemical
    modifications  and 13C  CP-MAS NMR spectroscopy.
    Environmental Science and Technology 38:5059-5063
    (2004).
Smolyakov, B.S., M.V.Zhigula, A.P Ryzhikh, E.V.Sinitsyna,
    N.I. Ermolaeva, and A.A. Fedotova. Copper(ll) specia-
    tion in a freshwater ecosystem.  Water Resources 31:
    55-63 (2004).
Steed, V.S., M.T. Suidan,  M. Gupta,  T. Miyahara, C.M.
    Acheson, and G.D. Sayles. Development of a sulfate-
    reducing biological process to remove heavy metals
    from acid mine drainage. Water Environment Research
    72: 530-535 (2000).
Stumm, W. and K. Morgan. Aquatic Chemistry, 3rd Edition.
    Wiley-lnterscience, New York, NY (1996).
Tabak, H.H., R. Scharp, J. Burckle, F.K. Kawahara, and R.
    Govind. Advances in biotreatment of acid mine drain-
    age and biorecovery of  metals:  1. Metal  precipitation
    for recovery and recycle. Biodegradation 14: 423-436
    (2003).
Tessier, A., P.G.C. Campbell, and M. Bison.  Sequential ex-
    traction procedure for the speciation of particulate trace
    metals. Analytical Chemistry 51: 844-851 (1979).
Thompson, R.A. and G.R. Helz. Copper speciation in sul-
    fidic solutions at low sulfur activity: Further evidence
    for cluster complexes?  Geochimica et Cosmochimica
    Acta 58: 2971 -2983 (1994).
USEPA. 2006 Edition of the Drinking Water Standards and
    Health Advisories,  EPA 822-R-06-013, Office of Water,
    Washington DC, (2006a).
USEPA. National Recommended Water Quality  Criteria,
    4304T, Office of Water, Washington DC, (2006b).
Waybrandt, K.R., D.W. Blowes, and C.J. Ptacek. Selection
    of reactive mixtures for use in permeable reactive walls
    for treatment of mine drainage. Environmental Science
    and Technology 32: 1972-1979 (1998).
Wedepohl,  K.H.   The composition  of  continental  crust.
    Geochimica  et Cosmochimica  Acta 59:  1217-1232
    (1995).
Wilkin, R.T. and M.S.  McNeil.  Laboratory evaluation of
    zero-valent iron to treat  water impacted by acid mine
    drainage. Chemosphere 53: 715-725 (2003).
Woinarski, A.Z., I. Snape, G. W. Stevens, and S.C. Stark.
    The effects of cold temperature on copper ion exchange
    by nature zeolite for use in a permeable reactive barrier
    in Antarctica.  Cold Regions Science and Technology
    37: 159-168(2003).
                                                    40

-------
Xiao, Z., C.H. Gammons, and A.E. Williams-Jones.  Ex-
    perimental study of copper(l) chloride complexing in
    hydrothermal solutions at 40 to 300 C and saturated
    water vapor pressure. Geochimica et Cosmochimica
    Acta 62: 2949-2964 (1998).
Zagury, G.J., R. Samson, and L. Deschenes. Occurrence
    of metals in soil and ground water near chromated cop-
    per arsenate-treated poles. Journal of Environmental
    Quality 32: 507-514 (2003).
                                                   41

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                                             Chromium
                         Douglas B. Kent, Robert W. Puls, Robert G. Ford
Occurrence and Distribution

 The average crustal abundance of chromium (Cr) is ap-
proximately 100 micrograms per gram. Chromium is much
less abundant than vanadium and manganese (Mn), which
are its nearest neighbors on the periodic table, but more
abundant than the other first-row transition elements cobalt
(Co),  nickel, copper, and zinc.  Chromium is a  group VI
transition metal, whose name refers to the variety of col-
ors associated with its various oxidation states.  Only two
of these oxidation states are known to occur naturally. In
minerals that crystallized in the earth's interior Cr occurs
exclusively in the plus 3 oxidation state (Cr(lll)). In aquatic
systems at pH values above 2 and in the presence of oxy-
gen, Cr(lll) is thermodynamically unstable compared to the
plus 6 oxidation state (Cr(VI)) and, therefore, is subject to
oxidation to Cr(VI).

Minerals  containing Cr(VI)  are rare.  The best known is
the lead (Pb) chromate mineral crocoite  (PbCrO4), which
occurs as spectacular  red-orange crystals in the oxidized
regions of some Pb deposits.  It is from  this mineral that
Cr was first isolated and identified.  Several other Cr(VI)
minerals have been identified in evaporite deposits; those
from the  Atacama Desert  in South America account for
most of these minerals. These evaporite minerals include
chromate  (CrO42~) or dichromite (Cr2O72~) combined with
sodium (Na), potassium (K), calcium (Ca), or barium (Ba)
with varying amounts of sulfate or other anions.

Chromium(VI) has been found to occur naturally in ground
water underlying the arid Paradise Valley in Arizona, USA,
at concentrations as  high  as hundreds  to thousands of
micrograms per liter (ug L1) (ones to tens of micromoles
per liter, uM) (Robertson,  1975). The ground water was
oxic, had alkaline pH values, moderate concentrations of
dissolved  salts, and was likely  very old.  More  recently,
Cr(VI) has been detected  in ground water with similar
chemical characteristics underlying arid or semi-arid basins
in California. As discussed further below, oxic ground water
with neutral-to-alkaline pH  values and moderate-to-high
concentrations of other anions, such as sulfate, possesses
the chemical conditions that favor the persistence of Cr(VI)
and promote its mobility.

Remedial Technologies

The goal of remediation schemes is to reduce the carcino-
genic, soluble, and mobile Cr(VI) to the less toxic and less
mobile Cr(lll), which forms minimally soluble precipitates.
Successful treatment of Cr(VI) hinges upon the formation
and stability of Cr(lll) precipitates. Ex-situ treatment tech-
nologies for ground water commonly use pump-and-treat
approaches with chemical reduction of the Cr(VI) to Cr(lll)
followed by precipitation.  In-situ technologies  currently
used for remediation of Cr(VI) contamination employ some
form of chemical reduction and fixation (e.g., geochemical
fixation, permeable reactive barriers (PRBs), and reactive
zones established through chemical injections).  There is
fairly extensive performance data available to evaluate the
potential for use of PRBs constructed using zerovalent iron
(Wilkin and Puls, 2003).

Regulatory Aspects

Cr(VI) is considered teratogenic (Abbasi and Soni,  1984),
mutagenic (Paschin et al., 1983) and carcinogenic (Ono,
1988).  According to the International Agency for Research
on Cancer (IARC), Cr(VI) is considered a powerful car-
cinogen and its presence in waters is cause for concern.
The national primary drinking  water standard set by the
USEPA for total Cr is 0.1  mg  L1 (USEPA, 2006a).  The
state of California maximum contaminant level in water is
50 ug L1.  No separate drinking water standard for Cr(VI)
has been established, but a separate standard has been
discussed for some time.  For non-potable water sources,
ambient water quality criteria (AWQC) that are protective
of  aquatic life may  serve as  alternative cleanup goals.
For Cr(VI), current statutes list  both acute and chronic
criteria for fresh waters as 0.016  mg L1  and 0.011 mg L1,
respectively, for a water hardness of  100 mg L1  (USEPA,
2006b). Adjustments to these criteria are to be applied for
waters with different hardness. An example of where this
criterion may  apply is a site where contaminated ground
water discharges to surface water.

Geochemistry and Attenuation Processes

Aqueous Spec/at/on

Aqueous speciation of Cr(VI)  varies with pH and Cr(VI)
concentration (Palmer and Puls, 1994). The dominant
aqueous species  above pH 6 is CrO42~. The dominant
species between pH 0 and 6 is HCrCy at Cr(VI) concen-
trations below approximately 0.003 moles L1 (160 mg L1)
and Cr2O72- at higher Cr(VI) concentrations. The species
H2CrO4 is  dominant only at pH < 0 (i.e., under extremely
acidic conditions). In the presence of Ca or iron in the plus
                                                   43

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3 oxidation state (Fe(lll)), solubility studies  indicate the
formation of CaCrO4 and FeCrO4" as solution complexes
(Perkins and Palmer, 2000; Baron and Palmer, 1996).

At near-neutral pH values, Cr(lll) concentrations in equi-
librium with Cr(OH)3 are less than 0.1 uM (approximately
5 ug L'1); concentrations increase with decreasing pH below
7 and increasing pH above 11  (Palmer and Puls, 1994). The
dominant solution species below pH 3.5 is Cr3+; at these low
pH values Cr(lll) can be mobile and appreciable concentra-
tions of Cr(lll) may be observed (Seaman et al., 1999). As
pH increases, the dominant species changes through a se-
quence of hydrolysis products: CrOH2+, Cr(OH)2+, Cr(OH)3,
Cr(OH)4- (Rai et al.,  1987). Oligomers like Cr2(OH)24+ and
Cr3(OH)45+ may form as metastable species at mildly acidic
to near-neutral pH values and moderate Cr(lll) concentra-
tions (Baes and Mesmer, 1976).

Cr(lll) forms strong complexes with a variety of organic
ligands. Once formed, slow exchange of inner-coordination-
sphere ligands around Cr(lll) may result in the persistence
of Cr(lll)-organic complexes  under conditions where they
are thermodynamically  unstable  This may  account for
reports of elevated concentrations of Cr(lll) at  near-neutral
pH values and high concentrations of organics in a Cr-con-
taminated wetland (Mattuck and Nikolaidis, 1996), tannery
effluent (Walsh and O'Halloran, 1996), and organic soil or
landfill leachates (Li and Xue, 2001).

Solubility

Chromium(VI) forms soluble compounds with the alkali
(Na, K) and alkaline earth (Mg, Ca) metal cations that are
typically present at the highest concentrations in  ground
water.  It does form a sparingly soluble salt with barium Ba,
which is ubiquitous in ground water but usually present at
trace concentrations; the mineral hashemite (Ba(Cr,S)O4) is
known from at least one location. The importance of BaCrO4
at sites with Cr(VI)-contamination has been suggested by
the agreement between ground-water Cr(VI) concentrations
and those calculated based on BaCrO4 solubility (Baron and
Palmer, 1996). Equilibrium computations suggest that a
significant fraction of Cr(VI) could be present as BaCrO4 at
neutral and alkaline pH values if there is an adequate source
of Ba (Figure 5.1). Barite (BaSO4) was used as the source
of Ba used in the computations in Figure 5.1; Ba concentra-
tions in equilibrium with barite are independent of pH in the
range shown in Figure 5.1. Another potential source of Ba
in ground water is the dissolution of aluminosilicate minerals.
These dissolution reactions are typically acid-catalyzed and,
therefore, the mass of  Ba available from these  reactions
should increase with decreasing  pH.  Chromium(VI) also
forms sparingly soluble salts  with a variety of  heavy metal
ions, of which Pb (PbCrO4) is the most important by virtue
of the fact that it has been identified at Cr(VI)-contaminated
sites (Palmer and Wittbrodt, 1991).

Other Cr(VI) compounds can form at very low or very high
pH values.  A laboratory experimental study showed the
formation of Fe(lll) hydroxy chromates mixed with hydrous
ferric oxide in the pH  range 1.5 to 3.5 (Olazabal et al., 1997).
KFe3(CrO4)2(OH)6, a chromate analogue to jarosite,  has
been identified in acidic, Cr(VI)-contaminated soil (Baron
et al., 1996). The results of equilibrium calculations, based
on thermodynamic data for Cr(VI) species and solids from
Baron and Palmer (1998) suggest that this solid is stable
below pH 3 (Figure 5.2). Formation of KFe3(CrO4)2(OH)6 in
the equilibrium computations was driven by high concen-
trations of Fe(lll), which, in turn, were driven by increased
solubility of soil  hydrous ferric oxide with decreasing  pH.
The predicted distribution of Cr(VI) between the aqueous
and solid phases at low pH is sensitive to the values as-
sumed for the solubility of hydrous ferric oxide and the total
concentrations of K and Cr(VI).

A Cr(VI) analogue of the  Ca, aluminum (Al), carbonate
mineral ettringite (viz., Ca6[AI(OH)6]2(CrO4)326H2O) was
observed in Cr(VI)-contaminated concrete, as was a nearly
pure Cr(VI) hydrocalumite (viz., 3[CaO][AI2O3][CaCrO4]nH
2O) (Palmer, 2000). Laboratory experiments showed that
Cr(VI)-ettringite  precipitated from Ca- and  Al-containing
solutions at pH values greater than 10 (Perkins and Palmer,
2000).  In systems where there is an excess of carbonate
over the available Ca, Cr(VI)-ettringite is thermodynamically
unstable with respect to calcite and gibbsite at high  pH.
Therefore, Cr(VI)-ettringite and related minerals are most
likely to be important in systems at high pH where there is
a large excess of Ca over carbonate or where precipitation
of calcium carbonate is inhibited.

Concentrations of Cr(lll) in aqueous  systems are limited
by the  low solubility of Cr(OH)3 (Rai et al., 1987).  Copre-
cipitation of Cr(lll) with hydrous ferric oxide drives Cr(lll)
concentrations lower and expands the pH  range over which
Cr(lll) concentrations are very  low (Palmer and Puls, 1994;
Sassand Rai, 1987).
  1.0
  0.8
=5 0.6
O
5
  0.4
  0.2
  0.0
                              Dissolved Cr(VI)
          HCr04-


Total Cr(VI) = 1 mM
Equilibrium with BaSO4
                                   Cr042-
                                        BaCrO4(s)
               Cr2072-
    2.0
          3.0
                4.0
                     5.0
                           6.0

                           PH
                                 7.0
                                       8.0
                                             9.0
                                                  10.0
Figure 5.1  Distribution of 1 mM (millimoles per liter,
            which equals approximately 52 mg L1)
            chromium(VI) plotted as a function of pH
            in equilibrium with barite. Thermodynamic
            data for aqueous Cr(VI) species from Baron
            and Palmer (1998) and for BaCrO4 and
            BaSO4 from Smith and Martell (1989).
                                                     44

-------
   1.0
8. o-8
8
^  0.6
   0.4
s
2
LJ-
   0.2
   0.0
              Aqueous Cr(VI)
                                     HCr0-
             Total Cr(VI) = 1 mM
             Total K = 1 mM
             Soil Fe(lll) solubility after Lindsay (1988)
         KFe3(Cr04)2(OH)6
    2.0
              3.0
                       4.0
                                5.0
                                          6.0
                                                   7.0
                            PH
Figure 5.2 Calculated distribution of 1mM Cr(VI) (ap-
            proximately 52 mg L1) in the presence
            of 1 mM K (approximately 39 mg L1) and hy-
            drous ferric oxide calculated as a function of
            pH. Thermodynamic data for aqueous and
            solid phase Cr(VI) from Baron and Palmer
            (1998) and for soil hydrous ferric oxide solu-
            bility from Lindsay (1988).

Adsorption

Adsorption of  Cr(VI) results from chemical reactions be-
tween aqueous Cr(VI) species and sites at mineral surfaces.
Adsorption becomes more favorable with decreasing pH be-
cause many of the Cr(VI) adsorption reactions consume H+
and the electrostatic  contribution to adsorption onto some
important adsorbents becomes more favorable at lower pH
values (Dzombak and Morel, 1990).  The intensity of bind-
ing of Cr(VI) at the mineral-water interface is intermediate
between those of strongly binding anions, like phosphate
and arsenate, and weakly binding anions like sulfate (Davis
and Kent, 1990). Asa result, other anions may out-compete
Cr(VI) for adsorption sites depending on  their concentra-
tions relative to Cr(VI) and the pH. Such competitive effects
can give rise to complex dependencies of adsorption on
pH and solute concentrations. This effect is illustrated  in
Figure 5.3. Adsorption of  Cr(VI) onto freshly precipitated
hydrous ferric  oxide (HFO) was determined over a range
of pH values and in  the presence of different concentra-
tions of sulfate (Leckie  et al., 1984). In  the  absence  of
sulfate, dissolved concentrations of  Cr(VI) decrease  with
decreasing pH and,  therefore,  so should the mobility  of
Cr(VI). As concentrations  of sulfate increase, competition
for adsorption sites  shifts in favor of sulfate and  Cr(VI)
adsorbs  less extensively. Thus, Cr(VI) mobility increases
with increasing concentrations of sulfate and, at constant
sulfate, it first  decreases then increases  with decreasing
pH (Figure  5.3).   Other  naturally occurring oxyanions,
such as carbonate and  silicate, also  compete with  Cr(VI)
for adsorption  sites (van Geen et al., 1994; Zachara et al.,
1987). In Cr(VI)-contaminated ground water,  competitive
adsorption with other anions, such as arsenate, phosphate,
vanadate (Leckie et al., 1984), or organic acids (Mesuere
and Fish, 1992) may produce similar effects. The influence
of pH and anion  concentration on mobility of  Cr(VI) has
been demonstrated in field-scale transport experiments in a
sand and gravel aquifer whose adsorption properties were
dominated by hydrous Fe- and Al-oxide coatings on quartz
and feldspar grains (Davis et al., 2000; Kent et al., 1995,
1994). Thus, the  mobility of Cr(VI) should  be expected to
vary with aquifer chemistry.

The affinity of soils and sediments for adsorption of Cr(VI)
varies widely depending on composition. Laboratory col-
umn experiments  have shown that adsorption of Cr(VI) dur-
ing transport is greatly enhanced by increasing abundance
of hydrous ferric oxide in sediments (Martin and Kempton,
2000; Stollenwerk and Grove, 1985).  Adsorption of Cr(VI)
onto aquifer sediment whose grain surfaces were coated
with Fe- and Al-containing hydrous oxides was significantly
less extensive than expected  from adsorption onto pure
hydrous oxides of Fe or Al (Kent et  al., 1995).  Adsorp-
tion of Cr(VI) over a range of pH values onto four different
soils varied from essentially  no adsorption to extensive
adsorption (Zachara et al., 1989); all four soils had high
contents of clay-sized minerals and high specific surface
areas. Results of these studies show that Cr(VI) adsorp-
tion onto soils or sediments cannot be predicted a priori
and, therefore, must be determined experimentally using
site-specific materials.
  5.0
  4.0

  1.0
  o.o
             Fe: 1 mM
             Total Cr(VI): 5 uM
                              No Sulfate
                                                0.25
                                                0.20
                                                0.15 c1
                                                0.10
                                                0.05
    4.0
            5.0
                     6.0
                             7.0
                                      8.0
                                              9.0
                                                0.00
                         PH
Figure 5.3 Concentration of dissolved Cr(VI) in equilib-
            rium with Cr(VI) adsorbed on freshly precipi-
            tated hydrous ferric oxide. The Fe concen-
            tration of 1 mM corresponds to approxi-
            mately 90 mg FeOOH/L The 5 u/W Cr(VI)
            concentration corresponds to approximately
            260 \jg/L.  Sulfate concentrations of 1 mM
            and 10 mM correspond to 96 and 960 mg
            SO/L, respectively  Continuous curves cal-
            culated from  the model for adsorption onto
            hydrous ferric oxide of Dzombak and Morel
            (1990). Experimental data from Leckie et al.
            (1986).
                                                      45

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Redox Chemistry

Reduction of Cr(VI)  to  Cr(lll) is a  potentially important
mechanism for natural attenuation of Cr in ground water.
Research on this topic has contributed to identifying poten-
tial Cr(VI) reductants and identifying how the rate and extent
of reaction varies with changes in solution and other condi-
tions. Reduction of Cr(VI) in aqueous solution is promoted
by low pH and dissolved organic carbon (Stollenwerk and
Grove, 1985). Dissolved Fe(ll) reduces Cr(VI) rapidly over
a wide range of solution conditions (e.g., Buerge and Hug,
1997; Eary and Rai, 1988). The rate of reduction by aque-
ous Fe(ll) decreases with decreasing temperature; the effect
of temperature is more pronounced at near-neutral than at
low pH (Sedlak and Chan, 1997).  Dissolved sulfide (e.g.,
Pettine et al., 1998),  sulfur-containing dissolved organic
compounds  (Schroeder and  Lee,  1975), and hydrogen
sulfide gas (Thornton and Amonette, 1999)  have been
shown to reduce Cr(VI). Aqueous  reductants like these
are not likely to  be of widespread importance in natural
attenuation because  of the limited extent to which mixing
occurs during transport; reductants associated with solids
are likely to be much more important in natural attenuation
of Cr(VI)  (Palmer and Puls, 1994).

Many compounds found in soils and  sediments have been
shown to be capable of reducing Cr(VI). Reduction of Cr(VI)
by Fe(ll) associated with surfaces of oxide minerals (e.g.,
magnetite, ilmenite, and Fe(ll)-substituted goethite, White
and Peterson, 1996; Bidoglioetal., 1993), silicates (IItonet
al., 1997;Earyand Rai, 1989),andsulfides(Pattersonetal.,
1997) has been reported.  Reduction by Fe(ll) associated
with minerals has been  observed in the presence of dis-
solved oxygen in laboratory experiments (llton et al., 1997)
and in field-scale transport experiments (Kent et al., 1994).
Laboratory experiments  have demonstrated that reduction
of Cr(VI) by soil humic and fulvic substances occurs at an
appreciable rate at pH 2 but the rate of reduction decreases
with increasing pH (Wittbrodt and Palmer, 1996), which is
consistent with reduction of Cr(VI) by synthetic, substituted
phenols reported by  Elovitz and Fish (1994). The rate of
reduction by soil humic  and fulvic substances decreases
with decreasing temperature in the range 15-55 C (Witt-
brodt and Palmer, 1996). The rate of reduction of Cr(VI) by
soil organic matter may decrease during the course of the
reaction as a result of  disappearance of more reactive com-
ponents of the complex soil organic compounds (Wittbrodt
and Palmer, 1996). Soil organic matter may also promote
Cr(VI) reduction by enhancing  the  reductive dissolution
of Fe(lll)  to produce  Fe(ll) (Wittbrodt and Palmer, 1996).
Reduction of Cr(VI) by low molecular weight aliphatic and
aromatic organic acids is catalyzed by various titanium-, AI-,
and Fe(lll)-oxides (Deng and Stone, 1996a, b) representa-
tive of solids found in soils and sediments.

The reactivity of Fe(ll) at mineral surfaces is subject  to
passivation as a  result of build up of oxidation products at
the surface. Reduction of Cr(VI) by magnetite and illmenite
was inhibited by build-up of Fe(lll) oxides on the surface
resulting from reaction between  Fe(ll)-containing minerals
and  Cr(VI) or by prolonged exposure of  Fe(ll) mineral  to
oxygen (Peterson et al., 1997; White, and Peterson, 1996).
Magnetite collected from anoxic reservoir sediments re-
duced Cr(VI) rapidly but magnetite collected from an oxic
soil profile did not reduce Cr(VI), presumably as a result of
passivation by Fe(lll) oxide coatings on the surface (White
and Peterson, 1996).  Passive oxide surface coatings can
adsorb Cr(VI) but can  completely inhibit its reduction (Pe-
terson etal.,  1996).

Oxidation of Cr(lll) to Cr(VI) is an important process to con-
sider for the long-term performance of a natural attenuation
alternative. Oxidation of Cr(lll) to Cr(VI) has been observed
in some soil and sediment slurries under oxidizing condi-
tions (Palmer and  Puls, 1994;  Masscheleyn et al.,  1992;
Bartlett and James, 1979). These studies involved adding
Cr(lll)-containing solutions  to  soil or sediment slurries.
Oxidation of naturally occurring Cr(lll) to Cr(VI) in soils has
been reported (Chung et al., 2001), but  the form in which
Cr(lll) occurred in the soils and the mechanism(s) by which
it was oxidized were unknown.

Laboratory experimental studies have provided insight into
possible  mechanisms  by which Cr(lll) can be oxidized to
Cr(VI) in soils and  sediments.  Oxidation of Cr(lll) by dis-
solved O2 in  homogeneous solution is too slow to be an
important process  at low temperatures (Nakayama et al.,
1981) but may be appreciable at elevated temperatures
(Schroeder and Lee, 1975). The potential for solid surfaces
to  catalyze the oxidation of Cr(lll) by dissolved oxygen
has not  been adequately studied.  Various Mn(IV) and
Mn(lll) oxides can oxidize Cr(lll) to Cr(VI) rapidly (Fendorf
and Zasoski, 1992; Eary and  Rai, 1987; Manceau, and
Charlet,  1992). The reaction is inhibited by precipitation
of Cr(OH)3 on the Mn oxide surface (Banerjee and Nesbitt,
1999; Fendorf et al., 1992). At near-neutral pH values, the
oxidation is not inhibited by Co(ll) or Mn(ll) but is inhibited
by Al (Fendorf et al., 1993), which apparently blocks sites
of oxidation  either  by surface precipitation or competitive
adsorption. The oxidation reaction is also inhibited by dis-
solved organic compounds that form complexes with Cr(lll)
(Johnson and Xyla, 1991; Nakayama et al., 1981). These
inhibitory effects suggest that  oxidation requires contact
between aqueous Cr(lll) and reactive sites at the Mn oxide
surface.  Thus, while Mn oxides are common constituents
of soils and aquifer sediments and, therefore, potentially
important oxidants for Cr(lll), their ability to oxidize Cr(lll)
will need to be determined on site-specific materials over
ranges of aquifer chemistry that are  relevant to field  ap-
plications.

Colloidal Transport

There is evidence in  natural systems  of  association of
chromium with mobile colloidal  solids in surface water
(e.g., Stolpe  et al., 2005) and ground water (e.g., Jen-
sen and  Christensen, 1999).  In  general, these colloids
consist of organic macromolecules,  iron oxyhydroxides,
clay minerals, or sulfide minerals. Increased mobility of
chromium-bearing colloidal material may result either from
changes in the surface charge on colloids due to changes
in subsurface geochemistry (e.g., Grolimund and Borkovec,
                                                     46

-------
2006) or through deflocculation and resuspension of col-
loidal material through dissolution of cementing agents
within the aquifer matrix (e.g., Ryan and Gschwend, 1990;
Ryan and Gschwend, 1992).  Both processes would be
facilitated in aquifers impacted by organic contaminants
where microbial activity may be stimulated resulting in the
generation of reducing conditions and/or the production of
low molecular weight organic compounds that partition to
fine-grained sediments. Due to its particle reactivity, mobile
forms of Cr(lll)  may be attributed,  in part,  to association
with mobile colloids. The distance of colloidal transport
from an impacted zone is uncertain, since colloid stability
may change significantly during transport from reduced to
oxidized zones or due to encounter with new sediment sur-
faces in unimpacted zones. However, identification of this
transport mechanism within a chromium plume may play
an important role relative to the management of the source
zone or the establishment of a monitoring system appropri-
ate for determining the extent of colloidal transport.

Site Characterization

Overview
In most cases the reduction of Cr(VI) to Cr(lll) will be the
process most likely to provide long-term attenuation and
immobilization of Cr in contaminant plumes and, therefore,
site characterization should focus on assessing the poten-
tial effectiveness of this process.  As noted in the previous
section, Cr oxidation-reduction chemistry is characterized
                             by disequilibrium. Therefore, assessments based on equi-
                             librium assumptions, such as Eh or pe determinations, will
                             be of little use.  Instead, it  is recommended that the as-
                             sessment focus on: 1) identifying solid and solution species
                             capable of reducing Cr(VI),  2) quantifying the capacity for
                             Cr(VI) reduction, 3) determining the quantity of Cr(VI) that
                             could be released overtime,  4) assessing possible changes
                             in  ground water chemistry that could influence the mobi-
                             lization of soil- and sediment-bound Cr(VI), 5) assessing
                             the long-term fate of reductants given possible changes in
                             ground-water chemistry, and 6) assessing the potential for
                             re-oxidation of Cr(lll).

                             Assessment of Cr Aqueous Speciation
                             Identification  of the mobile form of  chromium within the
                             aquifer is important relative to the development of a  con-
                             ceptual site model of the process(es) controlling immobi-
                             lization of this contaminant. Chromium may be mobile as
                             a dissolved species or in association with mobile colloidal
                             matter.  Generally, the anionic Cr(VI) species is the most
                             mobile form of dissolved chromium  in ground water with
                             neutral to basic pH and measurable oxygen (Ball and Izbicki,
                             2004), although  Cr(lll)  bound to mobile forms of natural
                             organic matter may occur under certain conditions (Li and
                             Xue, 2001).  The distribution of these potential species in
                             ground water provides information relative to the dominant
                             process controlling attenuation. Parks et al. (2004) provide
                             a review of several published methods for determining chro-
                             mium speciation in water samples, including those similar to
Table 5.1  Natural attenuation and mobilization pathways for chromium.
  Attenuation Processes
   Mobilization Processes
            Characterization Approach
 Precipitation of metal
 chromates or precipita-
 tion of Cr(lll) oxyhydrox-
 ide or sulfide
Dissolution of metal chro-
mates due to change in pH;
dissolution of Cr(lll) oxyhy-
droxide due to acidification;
dissolution of Cr(lll) sulfide
due to shift from reducing to
oxidizing conditions.
Evaluation of Cr speciation in ground water. Deter-
mination of total Cr in the solid matrix and suspected
components in chromium-bearing mineral.  Evaluation
of mineral solubility relative to ground-water chemistry
and published solubility constants.  Determination of
Cr(VI) reductants  in ground water and aquifer solids.
 Co-precipitation of Cr
 as a trace component in
 oxyhydroxides or sulfides
 of iron or manganese
Dissolution of host oxyhydrox-
ide due to decrease in pH or
shift from oxidizing to reducing
conditions; dissolution of host
sulfide due to shift from reduc-
ing to oxidizing conditions.
Evaluation of Cr speciation in ground water. Evaluation
of host precipitate formation relative to existing ground-
water chemistry; determination of host mineral content
in aquifer sediments via mineralogical characterization.
Evaluation of Cr solid-phase partitioning using sequen-
tial extraction methodologies.
 Adsorption of chromate
 to iron oxyhydroxides,
 iron sulfides, or other
 mineral surfaces
Desorption at high pH for
oxyhydroxides and sulfides;
complexation/stabilization in
the presence of DOC. Reduc-
tive dissolution of iron hydrox-
ides or oxidative dissolution of
iron sulfides.
Evaluation of Cr speciation in the aqueous phase.  De-
termination of total Cr in the solid matrix.  Evaluation of
Cr solid-phase partitioning using sequential extraction
methodologies. Batch and column testing to determine
Cr uptake capacity of site-specific aquifer materials
with variable geochemical conditions.
                                                     47

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published EPA methods (Table 5.2). These authors stress
the importance of sample preservation and processing on
the determination of total Cr and Cr(VI) in water samples,
particularly in samples  that may contain elevated con-
centrations of colloidal iron oxyhydroxides that may pass
through a 0.45 urn filter or precipitate in-situ during storage
of improperly preserved samples.

Differentiation of  Cr(lll)  and  Cr(VI) species  is  typically
determined via two approaches: 1) chemical separation
of the two species by ion-exchange or coprecipitation fol-
lowed by direct determination  of total dissolved chromium
and 2) selective determination of Cr(VI) by colorimetry or
differential pulse polarography. Colorimetric  methods for
direct determination of Cr(VI) exist for field measurements
and may  be employed if there are concerns about the abil-
ity to preserve chromium speciation prior to analysis in a
laboratory setting.  A recent study provides a useful com-
parison of the relative performance of  a field colorimetric
method, a field method for species separation followed by
laboratory analysis, and  EPA Method 218.6 for laboratory
separation and quantification  of Cr(lll) and Cr(VI)  by ion
chromatography (e.g., Ball and  McCleskey, 2003a and
2003b).  Ultimately, the  method employed for speciation
of aqueous chromium will depend on  required analytical
detection limits and sample preservation requirements, so
it is recommended that analytical  performance be demon-
strated on a site-specific basis.
It has also been observed that chromium may be leached
from certain grades of stainless steel well casing/screen
materials under chemical conditions that may be encoun-
tered in contaminant plumes.  Two published studies have
provided  detailed evaluation  of the extent of chromium
(and nickel) leaching that may occur for type 304 or 316
stainless steel screens (Hewitt, 1994; Oakley and Korte,
1996). Oakley and Korte (1996) provide a site-specific
example of how elevated chromium concentrations derived
from continuous leaching of well screen materials may be
falsely identified as a component of a ground-water plume.
This suggests that careful consideration should be given to
the types of well screen  materials and sampling protocols
employed relative to the  assessment of potential contami-
nants of concern within a plume.

Identifying and Quantifying Cr(VI)
       Reductants

Ground-water characteristics of direct interest to assess-
ment of Cr(VI) reduction potential include pH and concen-
trations of dissolved oxygen,  Fe(ll), sulfide, and organic
carbon. Ground-water pH can influence Cr(VI) reduction in
two ways. First, from a thermodynamic perspective, the af-
finity (Morgan, 1967) of many oxidation-reduction reactions
involving the Cr(VI)-Cr(lll) couple depends on pH. Second,
many of the rates of oxidation-reduction reactions depend
on pH. The presence of high concentrations of dissolved
Table 5.2  Published USEPA methods for determination of total chromium and speciation in aqueous samples.
Method Name and Number
Chromium, Hexavalent - Chelation Extraction;
7797
Chromium, Hexavalent - Colorimetric Method;
7196A
Chromium, Hexavalent - Coprecipitation Method;
7795
Chromium, Hexavalent - Differential Pulse Polarog-
raphy; 7798
Determination of Hexavalent Chromium in Drink-
ing Water, Groundwater, and Industrial Wastewater
Effluents by Ion Chromatography, 7799
Metals and Trace Elements by ICP/ Atomic Emis-
sion Spectrometry; 200.7 Rev. 4.4
Trace Elements by ICP/Mass Spectrometry; 200.8
Rev. 5.4
Trace Elements by Stabilized Temperature Graph-
ite Furnace AA Spectrometry; 200.9 Rev. 2.2
Source
a SW-846 Ch 3.3;
http://www.epa.gov/epaoswer/hazwaste/test/pdfs/7197.pdf
a SW-846 Ch 3.3;
http://www.epa.gov/epaoswer/hazwaste/test/pdfs/7196a.pdf
a SW-846 Ch 3.3;
http://www.epa.gov/epaoswer/hazwaste/test/pdfs/7195.pdf
a SW-846 Ch 3.3
htt p ://www. epa . gov/sw-846/pdf s/7 1 98 . pdf
a SW-846 Ch 3.3;
http://www.epa.gov/sw-846/pdfs/7199.pdf
b EPA/600/R-94/1 1 1 (NTIS Order Number PB95-1 25472)
b EPA/600/R-94/1 1 1 (NTIS Order Number PB95-1 25472)
b EPA/600/R-94/1 1 1 (NTIS Order Number PB95-1 25472)
a EPA SW-846 Ch 3.3, Test Methods for Evaluating Solid Waste, Physical/Chemical Methods
  (http://www.epa.gov/epaoswer/hazwaste/test/main.htm)
b Methods for the Determination of Metals in Environmental Samples-Supplement I, EPA/600/R-94/111, May 1994. Available at
  National Technical Information Service, PB95-125472. (http://www.ntis.gov/products/types/publications.asp?loc=4-4-4)
                                                     48

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oxygen  is indicative of conditions unfavorable for Cr(VI)
reduction.  In contrast, low but detectable dissolved oxy-
gen concentrations can indicate mildly reducing conditions
favorable for the accumulation of Cr(VI)-reductants in the
sediments. Detectable concentrations of Fe(ll) or sulfide
are indicative of conditions under which Cr(VI) reduction
should occur rapidly.   High concentrations of dissolved
organic  carbon could  indicate that organic compounds
capable of reducing Cr(VI), either directly or catalyzed by
mineral surfaces, may be present.

Detection of  Cr(VI) reductants  in  ground water signals
conditions favorable for natural attenuation,  but  most of
the Cr(VI) reduction capacity will likely reside  in the sedi-
ments (Palmer and Puls, 1994). Potentially important Cr(VI)
reductants include sulfides, Fe(ll) minerals, and organic
matter. Even in the absence of detectable concentrations of
reductants in ground water, Cr(VI) reductants may be pres-
ent in the sediments. For example, sediment-bound Fe(ll)
can provide a significant capacity for natural  attenuation
of Cr(VI) even in the absence of detectable concentrations
of Fe(ll) or sulfide and the presence of low but detectable
concentrations of dissolved oxygen (Kent et al., 1994).

In most cases the capacity for Cr(VI) reduction by soils or
aquifer sediments will have to be determined  experimen-
tally using site-specific materials. Laboratory experiments
can be used to assess the Cr(VI) reduction capacity and
provide insights into reduction rates by constituents of the
sediments. However, it should be noted that exposure to at-
mospheric oxygen can greatly decrease the Cr(VI) reductive
capacity of sediments (e.g., Anderson et al., 1994). Also,
rates of reduction measured in laboratory experiments may
exceed those achieved during transport.  For example, the
rate of reduction of Cr(VI) determined in laboratory batch
experiments greatly exceeded that during transport in a
sand and gravel aquifer because, in the field, the rate of
reduction was limited by the rate of  mass transfer across
sediment layers (Friedly et al., 1995).

Use of Cr Isotopes to Determine Reduction
      Rates

Recent work  suggests that in-situ Cr(VI) reduction rates
may be determined through the assessment of chromium
stable isotope ratios (53Cr/52Cr)  in ground water  (Ellis et
al., 2002).  Cr(VI) reduction causes  an enrichment of the
lighter isotopes in the reduced product at any given instant.
Faster reduction of lighter isotopes of Cr(VI) as compared
to heavier isotopes resulted  in measurable changes in
the isotopic composition of dissolved Cr(VI) along flow
paths where  reduction was occurring.  Thus, the extent
of Cr(VI) reduction could be assessed along a flow path
provided the  isotopic composition of the initial source is
known.  Additional  work indicates that sorption reactions
occurring  within the fringe of contaminant  plumes may
contribute to  the observed isotope fractionation,  but this
work also indicated that sorption would exert little impact
in more  mature  portions of the plume where sorption may
play a less dominant role (Ellis et al., 2004).  Theoretical
calculations support the mechanism of proposed isotopic
fractionation between Cr(VI) and Cr(lll) (Schauble et al.,
2004; Ottonello and Zuccolini, 2005a; Ottonello and  Zuc-
colini, 2005b) suggesting that characterization of chromium
stable isotope ratios may prove a useful tool for assessing
in-situ rates of Cr(VI) reduction.

Assessment of Cr Solid-Phase Speciation

Various chemical extraction procedures have  been pro-
posed for  quantifying soil-  and sediment-bound Cr(VI)
contamination.   Studies in which  the  effectiveness of
different  extraction techniques  were compared  showed
that  a method involving a  heated carbonate-hydroxide
solution was the most effective for recovering soluble and
insoluble forms of Cr(VI) (James et al., 1995; Vitale et al.,
1997). For sparingly soluble Cr(VI) salts, the method was
effective at quantifying PbCrO4 but only partially effective
at quantifying  BaCrO4, which may be an important  form
of Cr(VI) at some sites (Palmer and Puls,  1994).  This
method has been adopted by the USEPA (Method 3060A;
http://www.epa.gov/epaoswer/hazwaste/test/pdfs/3060a.
pdf) for determination of Cr(VI) in contaminated soils and
waste materials. It should be noted that Vitale et al. (1997)
demonstrated that this method may not be appropriate for
solid materials containing sulfide precipitates that reduce
Cr(VI) to Cr(lll) upon dissolution during extraction. Figure
5.1 suggests that a mildly acidic (pH 3-4) acid leach in the
presence of sufficient sulfate to precipitate  Ba  as BaSO4
might effectively dissolve BaCrO4,  but this  has not been
tested. Very low pH values should be avoided because of
the potential for causing Fe(lll) concentrations to increase
sufficiently to precipitate chromian jarosite (Figure 5.2). At
least one potential artifact that could result in overestimating
sediment-bound Cr(VI) has been noted. Storing soils so
that they are partially saturated (with water) may  result in
oxidation of native Cr(lll) to Cr(VI); drying the soils prior to
storage apparently eliminates this possible artifact (Chung
et al., 2001). Korolczuk (2000) has examined an adaptation
of the heated carbonate-hydroxide extraction method that
employs  voltammetric determination of  extracted Cr(VI)
in organic-rich soils containing predominantly Cr(lll)  solid
species.  This  approach was more reliable  for  Cr(VI) de-
tection than use of the colorimetric procedure  described
in Method 3060A.

Various extraction schemes have been employed to de-
termine the chemical speciation  of the  total  chromium
content of contaminated soils or sediments. The intent of
this solid-phase characterization approach is generally to
identify the solid-phase partitioning  reactions that govern
chromium mobility in the subsurface. However, as noted
by Szulczewski et al., (1997), there are limitations to the
selectivity of various solutions proposed to extract chromium
from labile and recalcitrant pools of solid species within
soils and sediments.  These authors as well as Wilkin  et al.
(2005) have made use of x-ray absorption spectroscopy to
assist in defining the chemical speciation of chromium in
contaminated soils and  aquifer sediments,  but little  work
has been conducted to date to better design and assess
the accuracy of chemical  extractions to determine the
chemical speciation of chromium in  solid matrices. Thus,
                                                     49

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currently Method 3060A provides the most thoroughly
tested approach for assessing the presence of Cr(VI)  in
soils and sediment.

 Long-Term Stability and Capacity
Changing aquifer chemical conditions will influence the
processes that  control the mobility and  natural attenua-
tion of Cr(VI). Changes in ground-water pH and dissolved
salt concentrations  can change the solubility of  Cr(VI)
compounds (Figures 5.1 and 5.2) and the extent to which
Cr(VI) adsorbs on the sediments (Figure 5.3). Acidification
of natural attenuation zones can cause the Cr(VI) reductive
capacity of the sediments to decrease (e.g., Anderson  et
al., 1994) or to increase (e.g., Wittbrodt and Palmer, 1996;
Stollenwerk Grove, 1985). Acidification can also increase
the mobility of Cr(lll) (e.g., Walter etal., 1994). Invasion  of
oxygen into reducing zones can result in oxidation of Cr(VI)-
reductants.  Establishment of completely oxic conditions
may result in production of Mn oxides or other compounds
capable of re-oxidizing Cr(lll)  previously sequestered  in
the sediments.
Laboratory experimental studies conducted during the site
assessment phase will provide insight into the impact  of
anticipated changes in aquifer chemistry on many of these
processes. However, the long-term performance of natural
attenuation under changing aquifer chemical conditions will
be difficult to predict with certainty. Therefore, long-term
monitoring of chemical conditions in and  around zones  in
which attenuation processes are active is recommended. It
is recommended that long-term monitoring include the criti-
cal water quality parameters  identified above. If chemical
conditions become increasingly oxidizing, then it is recom-
mended that long-term monitoring be supplemented with
periodic assessment of the  Cr(VI) reduction capacity  of
the sediments.

Tiered Analysis
Determination of the viability of chromium remediation  in
ground water via monitored natural attenuation will depend
upon proper assessment of contaminant loading to the
aquifer and prevailing geochemistry and mineralogy within
the contaminant plume and the down gradient zone prior
to the point(s) of compliance. The goal of  site assessment
is to demonstrate the process(es) controlling chromium
sequestration onto aquifer solids and the  long-term stabil-
ity of solid phase chromium as a function of existing and
anticipated ground-water chemistry. The following tiered
analysis structure for site characterization provides an ap-
proach to evaluate candidate sites and define the potential
limitations of MNA as  part of a remedy for ground-water
cleanup.
Tier I.  Site characterization under Tier I will  involve dem-
onstration that the plume is static or shrinking,  has not
reached  compliance boundaries,  and does not  impact
existing water supplies.  Once this is established through
ground-water characterization, evidence is  collected  to
demonstrate Cr partitioning  to aquifer solids within the
plume. Rapid movement of contaminants along preferred
flow paths in the unsaturated and saturated zones can by
induced by hydrologic events such as heavy rains  (e.g.,
McCarthy et al., 1998; Camobreco et al., 1996).  It will be
important to determine that such hydrogeologic features do
not result in contaminants bypassing zones where natural
attenuation is occurring. If natural attenuation processes
are active throughout the plume, then there should be an
observed increase in solid phase  concentrations  within
regions of the plume with higher aqueous concentrations,
e.g., near the source term.   This field partitioning data
may be  supplemented  by geochemical modeling  that
incorporates measured  water chemistry (e.g.,  pH, Eh,
and major ion chemistry) throughout the plume to assess
the potential for solubility control by metal chromates or
chromium oxyhydroxides or sulfides.  Since identification
of the chemical speciation of chromium in water samples
and aquifer sediments is critical towards determining the
attenuation mechanism(s), it is recommended that precau-
tions be taken to preserve chromium speciation during col-
lection, preservation, and processing of collected samples
(See recommendations and application of these  methods
in Wilkin  et  al., 2002 and Ford et al., 2005).  Identification
of active  sequestration to prevent chromium migration in
ground-water provides justification for proceeding to Tier
II characterization efforts.

Tier II. Under Tier II, the apparent rate and mechanism(s)
of attenuation are determined.  Estimates of a site attenu-
ation rate(s) can be assessed via a well transect along the
ground-water flow path. This analysis may be supplemented
by analysis  of the distribution of Cr isotopes in ground-water
along the flow path. In addition, time-series data may be
collected  at one or more monitoring points within the plume.
This information will allow assessment of the relative tim-
escales for  contaminant  immobilization and fluid transport
and determination of  whether remediation objectives can
be met within the required regulatory time frame. In addi-
tion, the mechanism(s) for attenuation need to be identified
under this stage of site characterization. This effort may
require determination of the chemical speciation of aque-
ous and solid phase chromium, which may be approached
according to the following scheme:

    1)  Determination of chromium solution speciation
       via  direct analytical measurements in combi-
       nation with speciation calculations based on
       characterized  ground-water chemistry (e.g.,
       Jensen and Christensen, 1999);
    2)  Calculation of the saturation state of ground
       water relative to measured aqueous chemistry
       (e.g., Wilkin et al., 2005) complimented by the
       possible isolation of discrete Cr mineral phases
       via  density separations (or other schemes) in
       regions of the aquifer with highest solid phase
       concentrations;
    3)  Determination of aquifer mineralogy to deter-
       mine the relative abundance of components
       with documented capacity for Cr sorption (e.g.,
       Amonette, 2002);
                                                     50

-------
    4)   Identification of chromium association(s) with
        the various solid phase components of aquifer
        solids through combination of chemical extrac-
        tions with microscopic/spectroscopic confirma-
        tion of phase associations, and;
    5)   Demonstration of concurrence between the site
        conceptual model and mathematical model(s)
        that describe arsenic removal mechanism(s).
It is recommended that identification of chromium chemical
speciation in aqueous and solid matrices be conducted us-
ing samples collected in a manner that preserves the in-situ
mineralogy and speciation of chromium.  The demonstra-
tion of concurrence between conceptual and mathematical
models describing arsenic transport will entail development
of site-specific parameterization of the chemical processes
controlling chromium solid-phase partitioning.

Tier III.  Once the partitioning mechanism(s) have been
identified for the site, the subsequent characterization ef-
fort under Tier III will involve determination of the stability
of immobilized Cr and the capacity of  the aquifer to sus-
tain continued uptake. The impact of potential hydrologic
changes, such as a shift in flow direction caused  by the
onset of pumping at nearby sites, needs to be determined.
A  well-constrained ground-water flow model of the site
should be capable of evaluating the impact of such changes.
It is recommended that the stability of immobilized Cr be
tested based on the anticipated evolution of ground-water
chemistry concurrent with plume shrinkage.  For example,
changes in ground-water redox potential due to oxygen in-
trusion or the occurrence of anaerobic microbial processes
can exert a significant influence on Cr partitioning to iron
sulfides  or iron oxyhydroxides, respectively. Therefore, it is
recommended that sediment leach tests be conducted to
characterize the magnitude of Cr mobilization as a function
of ground-water redox chemistry representative of existing
and anticipated site conditions.  This may involve the use of
microcosm tests that stimulate in-situ microbial populations
toward the development of redox conditions considered
deleterious for continued Cr immobilization.  It  is recom-
mended that the capacity for Cr uptake onto aquifer solids
be determined relative to the specific mechanism(s) identi-
fied in Tier II.  For example, if site characterization under
Tier II indicated that co-precipitation of  Cr with iron sulfide
occurs due to microbial degradation of organic compounds
coupled with sulfate reduction within the aquifer, it is rec-
ommended that the mass distribution  of organic carbon,
sulfate and ferrous iron to support this reaction  within the
aquifer be determined. This site-specific capacity can then
be compared to Cr mass loading within the plume in order
to assess the longevity of the natural attenuation process.
If site-specific tests demonstrate the stability of immobilized
Cr and sufficient capacity within the aquifer to sustain Cr
attenuation, then the  site characterization effort can prog-
ress to Tier IV.  For cases where contaminant stability is
sufficient but aquifer capacity is insufficient for capture of
the entire plume, then a determination of the benefits of
contaminant source reduction may be necessary.
Tier IV. Finally, under Tier IV a monitoring plan is estab-
lished along with contingency plans in the event of MNA
failure.  It is  recommended that the monitoring plan be
designed to establish both continued plume stability and to
identify changes in ground-water chemistry that may lead
to re-mobilization of attenuated Cr.  The specific chemical
parameters to be monitored will include those identified
under Tier III  that may halt Cr partitioning to aquifer sedi-
ments and/or result in solubilization of either discrete Cr pre-
cipitates or aquifer minerals that sequester Cr from ground
water.  For example, solution phase parameters that could
alter Cr(lll) precipitation include decreases in ground-water
pH. Similarly, increases in the concentration of competing
anions, such  as phosphate, could lead to re-mobilization
of adsorbed Cr(VI).  In contrast, the concentration of dis-
solved iron or sulfate may indicate the dissolution of an
important sorptive phase within the aquifer (e.g., reductive
dissolution of  iron oxyhydroxides or oxidative dissolution of
sulfides). Changes in these parameters may occur prior to
observed changes in solution Cr and, thus, serve as moni-
toring triggers for potential MNA failure. In this instance,
a contingency plan can be implemented that incorporates
strategies to arrest possible plume expansion beyond com-
pliance boundaries. Possible strategies to prevent plume
expansion include pump and treat operations, installation
of reactive  barriers  to enhance uptake capacity perpen-
dicular to the direction of plume advance, or enhancement
of natural attenuation processes within the aquifer through
the  injection of soluble reactive components.

References

Abbasi,  S.A. and  R.  Soni.  Tetratogenic  effects of
    chromium(VI) in the environment as evidenced by the
    impact  of  larvae of amphibian  Rand tigrina: Implica-
    tions in the environmental management of chromium.
    International Journal of Environmental  Studies 23:
    131-137(1984).
Anderson, L.D., D.B. Kent, and J.A. Davis. Batch experi-
    ments characterizing the  reduction of Cr(VI) using
    suboxic material from a mildly reducing sand and gravel
    aquifer.  Environmental Science and Technology 28:
    178-185(1994).
Amonette, J.E. Methods for determination of mineralogy and
    environmental availability.  In So/7 Mineralogy with En-
    vironmental Applications. J.B. Dixon and D.G. Schulze
    (Eds.), Soil Science Society of America Book Series;
    Soil Science Society of America,  Inc., Madison, Wl,
    pp. 153-197 (2002).
Baes, C.F and R.E. Mesmer. The Hydrolysis of Cations.
    John Wiley and  Sons, New York, NY, 489 p. (1976).
Ball, J.W. and R.B. McCleskey.  A  new cation-exchange
    method for accurate field speciation  of hexavalent
    chromium. USGS Water-Resources  Investigations
    Report,  2003-4018. (http://wwwbrr.cr.usgs.gov/proj-
    ects/GWC_chemtherm/pubs/WRIR%2003-4018.pdf)
    (2003a)
                                                     51

-------
Ball, J.W.  and R.B. McCleskey. A new cation-exchange
    method for accurate  field  speciation of  hexavalent
    chromium. Talanta 61: 305-313 (2003b).
Ball, J.W. and J.A. Izbicki. Occurrence of hexavalent chro-
    mium  in ground water in the western Mojave Desert,
    California. Applied Geochemistry 19: 1123-1135
    (2004).
Banerjee,  D.  and  H.W. Nesbitt. Oxidation of  aqueous
    Cr(lll)  at birnessite surfaces: Constraints on reaction
    mechanism. Geochimica  et Cosmochimica Acta  63:
    1671-1687(1999).
Baron, D. and C.D. Palmer. Solubility  of KFe3(CrO4)2(OH)6
    at 4 to 35 C.  Geochimica et Cosmochimica Acta 60:
    3815-3824(1996).
Baron, D. and C.D. Palmer. Solubility of KFe(CrO4)2-2H2O at
    4-75 C. Applied Geochemistry 13: 961 -973 (1998).
Baron, D., C.D. Palmer, and J.T. Stanley.  Identification of
    two iron-chromate precipitates in a Cr(VI)-contaminated
    soil. Environmental Science and Technology 30: 964-
    968(1996).
Bartlett, R. and B. James. Behavior of chromium in soils:
    III. Oxidation. Journal of Environmental Quality 8:  31-
    35(1979).
Bidoglio, G., PN. Gibson, M. O'Gorman, and K.J. Roberts.
    X-ray absorption spectroscopy investigation of surface
    redox  transformations of  thallium and chromium on
    colloidal mineral oxides. Geochimica et Cosmochimica
    Acta 57: 2389-2394 (1993).
Buerge, I.J. and S.J. Hug.  Kinetics and pH dependence
    of chromium(VI) reduction by iron(ll).  Environmental
    Science and Technology 31: 1426-1432 (1997).
Camobreco, V.J., B.K. RichardsT.S. Steenhuis, J.H. Peverly,
    and M.B. McBride. Movement of heavy metals through
    undisturbed and homogenized soil columns. So/7 Sci-
    ence 161: 740-750 (1996).
Chung, J.,  R.G. Burau, and R.J.Zasoski.  Chromate genera-
    tion by chromate depleted subsurface materials. Water,
    Air, and Soil Pollution 128: 407-417 (2001).
Davis, J.A. and D.B. Kent. Surface complexation modeling
    in aqueous geochemistry. In Mineral-Water Interface
    Geochemistry. M.F. Hochella and A.F. White  (Eds.),
    Mineralogical Society of America, Washington DC, pp.
    177-260(1990).
Davis,  J.A., D.B.  Kent, J.A. Coston,  K.M. Hess, and  J.L.
    Joye. Multispecies reactive tracer test in an aquifer with
    spatially variable chemical condition: Experimental  de-
    sign and observed tracer movement. Water Resources
    Research 36:  119-134 (2000).
Deng, B. and AT. Stone. Surface-catalyzed Cr(VI) reduction:
    the TiO2-Cr(VI)-mandelic acid system.  Environmental
    Science and Technology 30: 463-472 (1996a).
Deng,  B. and AT. Stone.  Surface-catalyzed Cr(VI)  re-
    duction: reactivity of different organic reductants and
    different oxide surfaces.  Environmental Science and
    Technology 30: 2484-2494 (1996b).
Dzombak, D.A. and  F.M.M. Morel.  Surface Complexation
    Modeling: Hydrous Ferric Oxide. John Wiley and Sons,
    New York, NY, 416 p. (1990).
Eary, I.E. and D. Rai.  Kinetics of chromium(lll) oxidation
    to chromium(VI) by reaction with manganese dioxide.
    Environmental Science and Technology 21:1187-1193
    (1987).
Eary, I.E. and D. Rai. Chromate removal from aqueous
    wastes by reduction with ferrous iron. Environmental
    Science and Technology 22: 972-977 (1988).
Eary, I.E. and Rai, D.  Kinetics of chromate reduction by
    ferrous ions derived from hematite and biotite at 25 C.
    American Journal of Science 289: 180-213 (1989).
Ellis, A.S., T.J. Johnson, and T.D. Bullen. Chromium isotopes
    and the fate of hexavalent chromium in the environment.
    Science 295: 2060-2062 (2002).
Ellis, A.S., T.M. Johnson, and T.D. Bullen. Using chromium
    stable isotope ratios to quantify Cr(VI) reduction: lack of
    sorption effects. Environmental Science and Technology
    38: 3604-3607 (2004).
Elovitz, M.S. and W. Fish. Redox interactions of Cr(VI) and
    substituted phenols: kinetic investigation. Environmental
    Science and Technology 28: 2161-2169 (1994).
Fendorf, S., M. Fendorf, D.L. Sparks, and R. Gronsky. Inhibi-
    tory mechanisms of Cr(lll) oxidation by 8-MnO2 Journal
    of Colloid and Interface Science 153: 37-54 (1992).
Fendorf, S.E. and R.J. Zasoski. Chromium(lll) oxidation by
    8-manganese oxide (MnO2) 1. Characterization. Envi-
    ronmental Science and Technology 26: 79-85 (1992).
Fendorf, S.E., R.J. Zasoski, and R.G. Burau. Competing
    metal  ion influences  on  chromium(lll) oxidation by
    birnessite. So/7 Science Society of America  Journal
    57: 1508-1515(1993).
Ford, R.G.  The Impact of Ground Water-Surface Water
    Interactions on Contaminant Transport with Application
    to an Arsenic Contaminated Site. EPA Environmental
    Research Brief, EPA/600/S-05/002.  Cincinnati, OH:
    U. S. Environmental Protection Agency.  (http://www.
    epa. gov/ada/down load/brief s/epa_600_s05_002. pdf)
    (2005).
Friedly, J.C., J.A. Davis, and D.B. Kent. Modeling hexavalent
    chromium reduction in groundwater in field-scale trans-
    port and laboratory batch experiment. Water Resources
    Research 31: 2783-2794 (1995).
Grolimund, D. and  M. Borkovec.  Release  of  colloidal
    particles in natural porous media by monovalent and
    divalent cations. Journal of Contaminant Hydrology 87:
    155-175(2006).
Hewitt, A.D. Dynamic study of common well screen ma-
    terials. Ground Water Monitoring and Remediation, 9:
    87-94(1994).
llton, E.S., D.R.Veblen, C.O. Moses, and S. P. Raeburn.The
    catalytic effect of sodium and lithium ions on  coupled
    sorption-reduction of chromate at the biotite edge-fluid
    interface.  Geochimica et Cosmochimica Acta 61(17):
    3543-3563(1997).
                                                    52

-------
James, B.R., J.C. Petura, R.J. Vitale, and G.R. Mussoline.
    Hexavalent  chromium extraction from soils: a com-
    parison of five  methods. Environmental Science and
    Technology 29: 2377-2381 (1995).
Jensen, D.L. and T.H. Christensen. Colloidal and dissolved
    metals in leachates from four Danish landfills. Water
    Research 33: 2139-2147 (1999).
Johnson, C.A. and A.G. Xyla.The oxidation of chromium(lll)
    to  chromium(VI) on the  surface of  manganite (y-
    MnOOH).   Geochimica et Cosmochimica  Acta  55:
    2861-2866(1991).
Kent, D.B., J.A. Davis, L.C.D.Anderson, and B.A. Rea.Trans-
    port of chromium and selenium in the suboxic zone of
    a shallow aquifer: role of adsorption processes. Water
    Resources Research 31: 1041 -1050 (1995).
Kent,  D.B.,  J.A. Davis,  L.C.D. Anderson, B.A. Rea, and
    T.D.Waite. Transport of chromium and selenium in the
    suboxic zone of a shallow aquifer: Influence of redox
    and adsorption  reactions. Water Resources Research
    30: 1099-1114(1994).
Korolczuk, M. Voltammetric determination of traces of Cr(VI)
    in the presence of Cr(lll)  and humic acid. Analytica
    ChimicaActa4-\4: 165-171 (2000).
Leckie, J.O., A.R. Appleton, N. B. Ball, K.F. Hayes, and B.D.
    Honey man. Adsorptive removal of trace elements from
    fly-ash pond effluents onto iron oxyhydroxide: Electric
    Power Research Institute Report, EPRI-RP-910-1. Palo
    Alto, CA, 1984.
Li, Y. and H. Xue. Determination of Cr(lll) and Cr(VI) species
    in natural waters by catalytic cathodic stripping voltam-
    metry. Analytica Chimica Acta 448: 121-134 (2001).
Lindsay, W.L. Solubility and redox equilibria of iron com-
    pounds in soils. In  Iron in Soils and Clay Minerals. J.W.
    Stucki, B.A.  Goodman, and U. Schwertmann (Eds.), D.
    Reidel Publishing Co., Boston, MA pp. 37-62 (1988).
Manceau, A. and L. Charlet. X-Ray absorption spectroscopic
    study of the sorption of Cr(lll) at the oxide-water inter-
    face 1. Molecular mechanism of Cr(lll) oxidation on Mn
    oxides.  Journal of Colloid and Interface Science 148:
    425-442 (1992).
Martin, T.A. and J.H. Kempton. In situ stabilization of metal-
    contaminated groundwater by hydrous ferric oxide: an
    experimental and modeling  investigation. Environmental
    Science and Technology 34: 3229-3234 (2000).
Masscheleyn, PH., J.H. Pardue, R.D. DeLauned, and W.H.J.
    Patrick. Chromium redox chemistry in a lower Mississip-
    pi Valley bottomland hardwood wetland. Environmental
    Science and Technology 26: 1217-1226 (1992).
Mattuck, R. and N. Nikolaidis. Chromium mobility in fresh-
    water wetlands: Journal of Contaminant Hydrology 23:
    213-232 (1996).
McCarthy, J.F., K.R. Czerwinski, W.E. Sanford, P.M. Jardine,
    and J.D. Marsh. Mobilization of transuranic radionu-
    clides from disposal trenches by natural organic matter.
    Journal of Contaminant Hydrology 30: 49-77 (1998).
Mesuere, K. and W. Fish. Chromate and oxalate adsorp-
    tion on  goethite. 2. Surface complexation modeling of
    competitive adsorption.  Environmental Science and
    Technology 26 (12): 2365-2370 (1992).
Morgan, J.J. Applications and limitations of chemical ther-
    modynamics in water systems. In Equilibrium Concepts
    in Natural Water Systems.  W. Stumm (Ed.), American
    Chemical Society, Washington DC, pp.1-29 (1967).
Nakayama, E., T. Kuwamoto, S. Tsurubo, and T Fujinaga.
    Chemical speciation of chromium in sea water Part 2.
    Effects  of manganese oxides and  reducible organic
    materials on the redox processes of chromium. Ana-
    lytica Chimica Acta 130: 401-404 (1981).
Oakley, D. and N.E. Korte. Nickel  and chromium in ground
    water samples as influenced by well construction and
    sampling methods. Ground Water Monitoring and Re-
    mediation, 16: 93-99 (1996).
Olazabal,  M.A., N.P  Nikolaidis, S.A.  Suib,  and  J.M.
    Madariaga.  Precipitation equilibria of the chromium(VI)/
    iron(lll)system and spectroscopic characterization of the
    precipitates.  Environmental  Science and Technology
    31:2898-2902 (1997).
Ono, B.I. Genetic approaches in the study of chromium tox-
    icity and resistance in yeast and bacteria In Chromium
    in the Natural and Human  Environments. J.  Nriagu
    and E. Nieboer (Eds.), John  Wiley, New York, NY, pp.
    351-368(1988).
Ottonello, G. and M.V. Zuccolini. Ab-initio structure, energy
    and stable Cr isotopes equilibrium fractionation of some
    geochemically relevant H-O-Cr-CI complexes. Geochi-
    mica et Cosmochimica Acta  69: 851-874 (2005a).
Ottonello, G. and M.V. Zuccolini. Ab initio thermochemistry
    of some geochemically relevant molecules in the sys-
    tem Cr-O-H-CI. Geochimica et Cosmochimica Acta 69:
    3505-3518 (2005b).
Palmer, C.D. Precipitates in a Cr(VI)-contaminated concrete.
    Environmental Science and Technology 34:4185-4192
    (2000).
Palmer, C.D. and R.W. Puls. Natural attenuation ofhexava-
    lent chromium in ground water and soils. Environmental
    Research Brief. Environmental Protection Agency, Ada,
    OK, 13 p. (1994).
Palmer,  C.D. and PR.  Wittbrodt. Processes affecting the
    remediation of chromium-contaminated sites. Environ-
    mental Health Perspectives 92: 25-40 (1991).
Parks, J.L., L. McNeill,  M. Frey, A.D. Eaton, A. Haghani, L.
    Ramirez, and M. Edwards.  Determination of total chro-
    mium in environmental water samples. Water Research
    38: 2827-2838 (2004).
Paschin, Y.V., V.I. Kozachenko, and L.E. Sal'nikova. Differ-
    ential mutagenic response at  the HGPRT locus in V-79
    and CHO cells after treatment with chromate. Mutation
    Research 122: 361-365 (1983).
Patterson, R.R., S. Fendorf, and M. Fendorf.   Reduction
    of hexavalent chromium by  amorphous iron sulfide.
                                                    53

-------
    Environmental Science and Technology 31:2039-2044
    (1997).
Perkins, R.B. and C.D. Palmer. Solubility of Ca?[AI(OH)6]2(
    CrO4)326H2O, the chromate analog of ettringite; 5-75
    C. Applied Geochemistry 15: 1203-1218 (2000).
Peterson, M.L., G.E. Brown, and G.A. Parks. Direct XAFS
    evidence  for heterogeneous redox reaction at the
    aqueous chromium/magnetite interface. Colloids and
    Surfaces A. Physicochemical and Engineering Aspects
    107:77-88(1996).
Peterson, M.L., A.F. White, G.E. Brown, and G.A. Parks. Sur-
    face passivation of magnetite by reaction with aqueous
    Cr(VI): XAFS and TEM results. Environmental Science
    and Technology 31:  1573-1576 (1997).
Pettine, M., I. Barra, L. Campanella, and F Millero. Effect of
    metals on the reduction of chromium(VI) with hydrogen
    sulfide. Water Research 32: 2807-2813 (1998).
Rai, D., B.M. Sass, and D.A. Moore.  Chromium(lll) hydroly-
    sis constants and solubility of chromium(lll) hydroxide.
    Inorganic Chemistry 26: 345-349 (1987).
Robertson, FN.  Hexavalent chromium in the ground  water
    in Paradise Valley, Arizona. Ground Water 13:516-527
    (1975).
Ryan, J.N. and P.M. Gschwend. Colloid mobilization in two
    Atlantic Coastal Plain aquifers: Field studies. Water
    Resources Research 26(2): 307-322 (1990).
Ryan, J.N. and P.M. Gschwend.  Effect of iron diagenesis
    on the transport of colloidal clay in an unconfined sand
    aquifer. Geochimica et Cosmochimica Acta 56:1507-
    1521 (1992).
Sass, B.M. and D.Rai. Solubility of amorphous chromium(lll)
    hydroxide solid solutions.  Inorganic Chemistry26 (14):
    2228-2232(1987).
Schauble, E., G.R. Rossman, and H.P.Taylor, Jr.Theoretical
    estimates of equilibrium  chromium-isotope fraction-
    ations. Chemical Geology 205: 99-114 (2004).
Schroeder, D.C. and G.F Lee. Potential transformations of
    chromium in natural waters. Water, Air, and Soil  Pollu-
    tion 4: 355-365 (1975).
Seaman, J.C., P.M. Bertsch, and L. Schwallie. In situ Cr(VI)
    reduction within coarse-textured, oxide-coated soil and
    aquifer systems using Fe(ll)  solutions. Environmental
    Science and Technology 33: 938-944 (1999).
Sedlak, D.L. and PG. Chan. Reduction of hexavalent chro-
    mium by ferrous iron. Geochimica et Cosmochimica
    A?fa61:2185-2192 (1997).
Smith, R.M. and A.E. Martell.  Critical Stability Constants,
    vol. 6: Plenum,  New York, NY, 643 p. (1989).
Stollenwerk, K.G. and D.B. Grove. Adsorption and desorp-
    tion of hexavalent chromium in an alluvial aquifer near
    Telluride, Colorado.  Journal of Environmental Quality
    14(1): 150-155(1985).
Stollenwerk, K.G. and D.B. Grove. Reduction  of hexavalent
    chromium in water samples acidified for preserva-
    tion.  Journal of Environmental Quality  14:  396-399
    (1985).
Stolpe, B., M. Hassellov, K. Andersson, and D.R. Turner.
    High resolution ICPMS as an on-line detector for flow
    field-flow fractionation; multi-element determination of
    colloidal size distributions in a natural water sample.
    Analytic Chimica Acta 535: 109-121 (2005).
Szulczewski, M.D., PA. Helmke, and W.F Bleam.  Com-
    parison of XANES analyses and extractions to de-
    termine  chromium speciation in contaminated soils.
    Environmental Science and Technology 31:2954-2959
    (1997).
Thornton, E.G. and J.E. Amonette. Hydrogen Sulfide Gas
    Treatment of Cr(VI)- contaminated sediment samples
    from a plating-waste disposal site-Implications for in-
    situ remediation. Environmental Science and Technol-
    ogy 33: 4096-4101 (1999).
USEPA. 2006 Edition of the Drinking Water Standards and
    Health Advisories, Office of Water, Washington DC,
    EPA 822-R-06-013 (2006a). http://www.epa.gov/water-
    science/criteria/drinking/dwstandards.pdf
USEPA. National Recommended Water Quality Criteria,
    Office of Water, Washington DC,  4304T (2006b). http://
    www.epa.gov/waterscience/criteria/nrwqc-2006.pdf
van Geen, A., A.P. Robertson, and J.O. Leckie.  Complex-
    ation of carbonate species  at the goethite  surface:
    Implications for adsorption of metal ions  in natural
    waters. Geochimica et Cosmochimica Acta 58: 2073-
    2086(1994).
Vitale,  R.J., G.R. Mussoline,  K.A. Rinehimer, J.C. Petura,
    and B. R. James. Extraction of sparingly soluble chro-
    mate from soils: Evaluation of  methods and Eh-pH
    effects.  Environmental Science and Technology 31:
    390-394(1997).
Walsh, A.R. and J. O'Halloran.  Chromium speciation in
    tannery  effluent II. Speciation in the effluent and in
    a receiving estuary.  Water Research 30: 2401-2412
    (1996).
Walter, A.L., E.G. Frind, D.W. Blowes, C.J. Ptacek, and J.W.
    Molson.  Modeling of multicomponent reactive transport
    in groundwater 2. Metal mobility in aquifers impacted
    by  acidic mine tailings discharge. Water Resources
    Research 30 (11): 3149-3158 (1994).
White, A.F. and M.L. Peterson. Reduction of aqueous tran-
    sition metal species on surfaces of Fe(ll)-containing
    oxides. Geochimica et Cosmochimica Acta 60: 3799-
    3814(1996).
Wilkin,  R.T., R.D.  Ludwig, and R.G. Ford.   Workshop on
    Monitoring Oxidation-Reduction Processes for Ground-
    water Restoration, EPA/600/R-02/002.   Cincinnati,
    OH: U.S. Office of Research and  Development, U.S.
    Environmental Protection Agency (2002).
Wilkin,  R.T. and R.W. Puls.  Capstone Report on the Ap-
    plication, Monitoring, and Performance of Permeable
    Reactive Barriers for Ground-Water  Remediation,
    Volume  1: Performance Evaluations  at Two  Sites,
    EPA/600-R-03-045a. Cincinnati,  OH: U.S. Office  of
                                                    54

-------
    Research and Development, U.S. Environmental Pro-
    tection Agency, (2003).
Wilkin, R.T., C. Su, R.G. Ford, and C.J. Paul. Chromium-re-
    moval processes during groundwater remediation by a
    zerovalent iron permeable reactive barrier. Environmen-
    tal Science and Technology 39: 4599-4605 (2005).
Wittbrodt,  PR. and C.D. Palmer. Effect of temperature,
    ionic strength, background electrolytes, and Fe(lll) on
    the reduction of hexavalent chromium by soil humic
    substances.  Environmental Science and Technology
    30:2470-2477(1996).
Zachara, J.M., C.C. Ainsworth, C.E. Cowen, and C.T. Resch.
    Adsorption of chromate by subsurface soil horizons.
    So/7 Science Society of America Journal 53: 418-428
    (1989).
Zachara, J.M., D.C. Girvin, R.L. Schmidt, and C.T. Resch.
    Chromate adsorption on amorphous iron oxyhydroxides
    in the presence of major groundwater ions. Environmen-
    tal Science and Technology 21: 589-594 (1987).
                                                   55

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                                               Arsenic
                        Robert G. Ford, Douglas B. Kent, Richard T. Wilkin
Occurrence and Distribution

The types of arsenic wastes at Superfund sites include
byproducts of wood preserving, arsenic waste disposal,
pesticide production and application, and mining (Table 1 in
USEPA, 1997). Mandal and Suzuki (2002) provide a review
of anthropogenic sources of arsenic in terrestrial systems.
The sources of arsenic in soils are diverse, and they may
be derived from organic or  inorganic forms.  For example,
Foster et al. (1998) examined the chemical speciation of
arsenic in mine wastes to  show  that  the waste material
was chemically complex and composed of a combination
of arsenic in native mineral forms and sorbed to common
weathering products.  Jackson and Bertsch (2001) exam-
ined  the types of aqueous chemical  species of arsenic
released from leaching of organo-arsenic materials used in
agricultural applications.  The aqueous species of arsenic
released during the weathering of arsenic compounds in
anthropogenic sources may be distinct from the subsequent
speciation of arsenic in ground water. Leached species
may  be unstable in the ground-water environment, but
serve as intermediates that control the overall release of
arsenic from source zones.

Arsenic in soil and ground water is derived  from natural
and/or anthropogenic sources (USEPA, 1997).   Natu-
ral sources of arsenic are  derived from a wide array of
geologic materials, including igneous, metamorphic and
sedimentary rocks (Korte and Fernando,  1991; Welch et
al., 2000; Smedley and Kinniburgh, 2002).  Arsenic may
subsequently be accumulated during  secondary mineral
formation in overburden  materials and soils. In contrast,
anthropogenic sources are  typically derived from the land
application of  arsenical pesticides and  herbicides and
from disposal of arsenic-bearing wastes generated during
processing of ore materials for production of commercial
products (USEPA, 1995).

Natural sources of arsenic  may also pose a health risk in
ground water due to changes  in site geochemistry that
promote mobilization. An example of  this scenario is the
mobilization of naturally occurring  arsenic during transport
of landfill leachate through the ground-water aquifer (Houn-
slow, 1980). In this case, arsenic is mobilized via desorption
induced by competitive interactions with dissolved constitu-
ents or due to dissolution of  host mineral phases as a result
of microbial degradation of  organic compounds within the
contaminant plume. In order to evaluate the potential for
down gradient attenuation of mobilized arsenic, it is recom-
mended that site characterization be sufficient to identify
the potential sinks for arsenic and the chemical conditions
that would result in uptake onto aquifer solids.

Plume Characteristics

Hounslow (1980) and Smedley and Kinniburgh (2002) pro-
vide an assessment of geochemical triggers that may lead
to arsenic mobilization in subsurface systems.  In general
these include: 1) desorption at  high pH  under oxidizing
conditions and/or due to the influx of dissolved ions that
compete for sorption sites on aquifer minerals, 2)  desorp-
tion/dissolution due to a change to a reducing chemical
environment, and 3) mineral dissolution. The first process is
a result of the influx of dissolved constituents that compete
for or displace arsenic adsorbed to mineral surfaces without
a concomitant change in sorbent stability. The second pro-
cess is a result of a change in the ground-water chemistry
to a condition  under which the sorbent material is no longer
stable. This may be a result of a change in redox, pH or
other  factors  leading  to dissolution of the sorbent phase.
The third process is a result of a shift in  arsenic mineral
stability due to a change in ground-water chemistry. In this
instance, arsenic may represent a major or minor compo-
nent within  the solid,  e.g., orpiment or arsenic-rich pyrite,
respectively.  While site-specific mechanisms  may differ
between naturally occurring versus contaminant  sources
of arsenic,  these general arsenic mobilization  processes
apply under both scenarios.

Remedial Technologies

Technologies  for the remediation of arsenic in soil involve
treatments to  effect containment, immobilization, or sepa-
ration/concentration within the solid matrix. The  separa-
tion/concentration process is followed by some secondary
immobilization treatment (USEPA, 1995;  USEPA, 1997).
Technologies  for the treatment of arsenic  in ground water
are based on ex-situ or in-situ approaches.  Pump-and-
treat technologies  make use of processes common to
water and wastewater treatment for removal of dissolved
arsenic (USEPA, 2002). In-situ treatment technologies are
less common, but there is emerging research based on
the application of permeable reactive barriers for arsenic
removal from  ground  water. This technology is based on
installation of  reactive solid material into the subsurface to
intercept and treat the contaminant plume (Lackovic et al.,
2000; Su and Puls, 2001; Cheng et al., 2005).
                                                    57

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Regulatory Aspects

Arsenic is  the second most  common  contaminant of
concern (COC) for sites listed on the Superfund National
Priorities List (USEPA, 2002).  Of the 1209 sites on the
National Priorities List for which a Record of Decision has
been signed, there are 380 and 372 sites with arsenic as a
COC in ground water and soil, respectively. The maximum
contaminant level for arsenic in drinking water was recently
revised from 0.05 mg L1 to  0.01  mg  L1 (USEPA, 2001;
USEPA, 2006a; http://www.epa.gov/waterscience/criteria/
drinking/dwstandards.pdf). This may impact site-specific
cleanup goals at hazardous waste sites for locations where
ground water is the primary source of potable  water. For
non-potable water sources, ambient water quality criteria
(AWQC) that  are  protective of  aquatic life  may serve as
alternative cleanup goals.  For arsenic, current statutes list
both acute and chronic criteria for arsenic in fresh waters as
0.34 mg L1 and 0.15 mg L1, respectively (USEPA, 2006b;
http://www.epa.gov/waterscience/criteria/nrwqc-2006.pdf).
Contaminant cleanup goals vary widely for  arsenic in soil
depending on the use of total or leachable  metals as the
cleanup criterion and the potential impact and intended use
of ground water supplies (USEPA, 2001).

Geochemistry and Attenuation Processes

Aqueous Spec/at/on

In oxidizing environments, the predominant form of arsenic
in solution is arsenate, HnAsO4n'3. The arsenate oxyanion
may be protonated to various  degrees as a  function of
pH, but it is commonly present  as the negatively-charged
H2AsO41" or HAsO42" within the pH range of natural waters
(Cullen and Reimer,  1989; Nordstrom and Archer, 2003).
However, the neutral (H3AsO4) or the fully deprotonated
form (AsO43") may exist at acid or alkaline pH extremes that
may occur at contaminated sites.

The speciation of arsenic may be more complex in  re-
ducing environments.  In general,  arsenite,  HnAsO3n'3,
is the  predominant  arsenic species, but  mononuclear
(HxAsOySz(3+x~2y~2z))  and polynuclear thioarsenic species
may also form in sulfate-reducing zones in which iron is
depleted (Clarke and Helz, 2000; Wilkin et al., 2003; Bostick
et al., 2005; Hollibaugh et al.,  2005).  The  distribution of
various reduced arsenic species under relevant geochemi-
cal conditions for an arsenic contaminated site  is shown in
Figure 6.1. These model results  suggest that arsenite would
predominate under most reducing environments. However,
thioarsenic species may become significant in sulfate-reduc-
ing  zones that are depleted in iron. These conditions may
be encountered in organic-rich  ground water derived from
contaminated sites (Vroblesky  and Chapelle, 1994). The
aqueous concentration of thioarsenic species may remain
high in sulfide-dominated  systems due to the solubility of
orpiment at circumneutral pH (Webster, 1990).  The detec-
tion of  the species controlling aqueous arsenic speciation
in reducing environments may be complicated by the need
to employ different sample preservation techniques in the
absence or presence of aqueous sulfide.
         -6
         -10
         -4-
         -6-
         -10
pH6
1 mg/LAs
                - arsenite
                - total thioarsenic
pH6
1 mg/LAs
5mg/L total F^
                          (C)
pH8
1 mg/LAs
             -7
                                            -3
                    .5  .4   .3  -7  -6  -5
                          Total HjS
Figure 6.1  The distribution of arsenite and thioarsenic
            species in a reducing ground water. The
            sum of all possible thioarsenic species is
            shown for simplicity.  The dashed line indi-
            cates the current MCL for arsenic in drinking
            water (10 ppb). Model results were gener-
            ated using MinteqA2 following modification
            of the database to include mononuclear
            thioarsenic equilibrium expressions (Wilkin
            et al., 2003). Calculations were carried
            out with a background electrolyte of 0.1 M
            NaNO3, and mackinawite (FeS) was allowed
            to precipitate in systems with Fe2+.

Solubility

The precipitation of pure phase arsenic minerals is not likely
in ground-water systems. Examination of solubility trends
for phases such as As2O5, As2O3,  and As2S3 indicates that
these phases would form only under extreme conditions.
Metal arsenates are  relatively insoluble (e.g., Ca3(AsO4)2
and  Ba3(AsO4)2), but formation of these phases is  limited
by the typically lower concentrations  of dissolved  metals
and arsenic encountered in ground water downgradient from
contaminant source areas. The stability fields of scorodite
[Fe(lll)AsO4-2(H2O)] and possible arsenic su If ides are shown
in Figure 6.2 as a function of pH and Eh for concentrations
of total dissolved As, Fe, and S that may be encountered
within a contaminant plume. There is also recent evidence
for the microbially-mediated formation of ferric/ferrous iron
arsenites under certain geochemical environments  (Morin
et al., 2003). Under reducing conditions, the stability fields
of pure arsenic sulfides are quite narrow for relatively high
total concentrations of arsenic in ground water, although
site-specific conditions may support precipitation of sulfides
(e.g., O'Day et al., 2004). In addition, while the precipitation
of scorodite seems feasible,  this process is typically  limited
by the consumption of dissolved iron during the formation
of iron oxides or iron sulfides.  Thus, the immobilization of
arsenic via precipitation of pure As-bearing mineral phases
is typically  limited.   However,  coprecipitation with other
                                                     58

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common soil/sediment minerals such as iron oxides and
iron sulfides is a likely sink for arsenic in ground water.  For
this process, arsenic uptake is concurrent with formation of
metastable phases of  Fe-bearing minerals that can occur
proximate to transitions from oxidizing-to-reducing condi-
tions (or vice versa) within the subsurface (e.g., Ford et
al., 2005; Ford et al., 2006; Wilkin and Ford, 2006). This
process  is generally distinct from sorption to  pre-exist-
ing mineral phases in that arsenic may subsequently be
transferred to more stable mineral phases over time (e.g.,
Moore et al., 1988; Ford, 2002; Lowers et al., 2007). Given
the uncertainty in the rates of transformation of metastable
minerals to more stable forms, it is currently not possible to
reliably estimate the importance of this stabilization process
in ground-water systems.  However, it is recommended that
this process be considered in the context of the dynamics
of site-specific geochemistry and the required timeframe
for attainment of remedial goals.

Adsorption

The primary forms of  inorganic arsenic in both  oxidizing
and reducing ground water are oxyanions orthiooxyanions
(Ferguson and Gavis,  1972; Wilkin et al., 2003; Bostick et
al., 2005). Adsorption of arsenic species at mineral surfaces
occurs as a result of a set of chemical reactions between
aqueous species and  surface sites (Dzombak and Morel
1990; Davis and Kent, 1990).  Adsorption of both As(lll) and
As(V) onto most minerals exhibits a strong pH dependence
because: 1) most adsorption reactions between As(lll) and
As(V) and mineral surface sites  have H+ as a reactant, 2)
arsenic speciation varies  with pH (Figure 6.2), and 3)  the
electrostatic contribution to the free energy of adsorption
Figure 6.2 Eh-pH diagram for arsenic at 25C. System
            As-H2O-Fe-S, with ZAs=1&5 (750 \jg L1),
            ZFe=10-4 (5.6 mg L1j, andZS=10-3
            (32 mg L1J.  Solid stability fields are shaded
            blue.  Orpiment(a) and Realgar(a) are poorly
            crystalline forms of As2S3 and AsS, respec-
            tively. Arsenopyrite and crystalline forms of
            As2S3 and AsS are suppressed.
of As species onto most minerals varies with pH.  In addi-
tion to pH, the extent to which As(lll) or As(V) adsorbs at
mineral surfaces will be influenced by the concentrations
of other anions, which compete for surface sites, and cat-
ions, the adsorption of which can influence the electrostatic
contribution to anion adsorption.

Arsenic sorption has been demonstrated for a  wide range
of minerals common to soils and sediments with iron oxides
and sulfides appearing to play a dominant role  in oxidizing
and  reducing  environments, respectively (Goldberg and
Glaubig, 1988; de Vitre et al., 1991; Morse, 1994; McNeill
and Edwards,  1997; Manning et al., 1998;  Chiu and Her-
ing, 2000; Wolthers et al., 2005). The relative  distribution
of inorganic species of arsenic in relation to the distribution
of potential Fe-bearing minerals as a function of  pH and
Eh in the absence of significant sulfate reduction is shown
in Figure 6.3.  For  near-neutral pH under oxidizing condi-
tions, sorption of arsenate onto iron oxides such as goethite
(oc-FeOOH) would likely predominate. Under iron-reducing
conditions, sorption of arsenate or arsenite  to reduced
Fe-bearing minerals such as siderite (FeCO3), magnetite
[Fe(lll)2Fe(ll)O4], or green rust phases [Fe(ll)gFe(lll)2(OH)18
 4(H2O)] may predominate. However, arsenic may remain
mobile in Fe-reducing systems in which reduced Fe-bearing
minerals do not precipitate, resulting in a loss  of sorption
capacity within the aquifer sediment (Swartz et al., 2005;
Polizzotto et al., 2005; Polizzotto et al., 2006). For reduced
systems in which  iron-  and sulfate-reduction processes
are significant, arsenic may sorb to iron  sulfides such as
pyrite (Figure  6.4). Sorption of arsenic  to iron (or other
metal sulfides) may be  the dominant mechanism for ar-
senic mobilization  under sulfate-reducing conditions that
are undersaturated with respect to precipitation of a pure
arsenic sulfide (Wilkin and Ford, 2006).

The extent to which inorganic arsenic will partition to mineral
surfaces will be governed by competition with other anions
in solution and the net surface charge that develops as a
function of site-specific geochemistry.  There  are several
commonly occurring anions  in natural waters that could
compete with arsenic sorption to mineral surfaces. These
competitive reactions will be active for all  arsenic aqueous
species  in oxidized and reduced systems.  There is evi-
dence to support that arsenic desorption may occur due
to competition with other inorganic and organic anions in
solution or due to increases  in ground-water pH (Peryea,
1991; Jackson and Miller, 2000; Meng et al.,  2000; Red-
man et al., 2002).

Redox Chemistry

Microbial interactions in aqueous systems can impact ar-
senic mobility via direct and indirect mechanisms.  Experi-
mental evidence suggests that the speciation of  arsenic can
be directly influenced via microbially-mediated reduction-
oxidation (redox) reactions (Ahmann et al., 1994; Zobrist
et al., 2000). Common outcomes of these interactions are
1) the conversion of inorganic arsenic between  its oxidized
or reduced oxyanionic forms and 2) methylation-demethyl-
ation of arsenic (Cullen and Reimer, 1989; Anderson and
Bruland, 1991; Maeda, 1994). The extent to which these
                                                     59

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       a)
      .
      LU
        -.5
b)
                                                       LU
                                                        -.5
                                                           25C
                                                                                            PO2=1 bar
          0     2     4     6     8     10    12    14     0     2     4      6      8     10    12    14
                              pH                                              pH
Figure 6.3  Eh-pH diagrams for arsenic and iron at 25C for iron-reducing systems. These paired diagrams show the
            relative distribution of potentially adsorbing arsenic species (left) relative to representative types of Fe-
            bearing sorbents (right) that are predicted to occur as a function of Eh and pH.  (a) System As-H2O, with
            Z/As contoured from  1Q-8 to 1&4; region for elemental arsenic is shaded gray, (b) System Fe-C-H2O (no
            sulfur) with ZFe contoured from 1&2 to 1&6 andZC=10'3; Hematite and Wustite are suppressed.
          a)
           -.5
b)
                                                      LU
                                                        -.5
                              6     8

                                PH
                                        10    12    14
                   6     8

                     PH
                                                                                     10    12     14
Figure 6.4  Eh-pH diagrams for arsenic and iron at 25C for coupled iron- and sulfate-reducing systems. These
            paired diagrams show the relative distribution of potentially adsorbing arsenic species (left) relative to
            representative types of Fe-bearing sorbents (right) that are predicted to occur as a function of Eh and pH.
            (a) System As-S-H2O, with Z/As=7i75 andZS=10'3; all solids suppressed to show stability fields for the
            aqueous species,  (b) System Fe-C-S-H2O with'LFe=10-4, ZC=7i73, and 1.8=10~3; Hematite is suppressed.
                                                     60

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processes influence on arsenic mobility in intermediate and
deep ground-water systems is not well known.  However,
these processes could play a more significant role in shallow
systems with zones of ground-water discharge-recharge
where biological activity may be more pronounced.  More
commonly, microbial interactions govern arsenic mobility
by controlling the redox chemistry of major elements with
which arsenic is associated, e.g., iron, sulfur, and carbon
(Moore, 1994; Harrington et  al.,  1998; McGeehan et al.,
1998; Jones et al., 2000).  The development of iron- and/or
sulfate-reducing zones in ground-water contaminant plumes
governs the precipitation-dissolution  of iron- and sulfur-
bearing minerals. These reactions influence arsenic mobility
through production or  loss of sorptive material within the
aquifer (Ford, 2005).

Recent  studies have  increased awareness of  the role
of microorganisms in  catalyzing the  oxidation  of As(lll).
Chemoautotrophic microorganisms that can oxidize As(lll)
to As(V) using oxygen have been isolated from arsenopy-
rite-bearing rock (Santini et al.,  2000). Chemoautotrophic
microorganisms that can oxidize As(lll) under anaerobic
conditions using nitrate, nitrite,  and, possibly, Fe(lll) have
also been identified (Oremland et al., 2002). Laboratory
experiments have shown that these microorganisms can
oxidize millimolar concentrations of As(lll) over a period of
several hours under both aerobic and anaerobic conditions.
Arsenic-oxidizing Chemoautotrophic anaerobes were shown
to be responsible for maintaining As(V) concentrations up
to 160 uM in the anoxic hypolimnion of a lake whose sedi-
ments were heavily contaminated with As from industrial
activities (Senn and Hemond, 2002). The dominance of
As(V) over As(lll) in other anoxic water bodies underscores
the potential importance of microbial As(lll) oxidation (Senn
and Hemond, 2002).

Microbial reduction of As(V) to As(lll) is likely to be important
in some systems.  Anaerobes whose dominant  metabolic
process involves coupling As(V) reduction to  oxidation
of organic compounds or hydrogen  have been isolated
from natural  and polluted water or  sediments  (Ahmann
et al., 1994; Hoeft et al., 2002).  Laboratory studies have
shown  that these organisms  can reduce As(V)  adsorbed
to hydrous ferric or aluminum oxides (Zobrist et  al., 2000)
or precipitated with Fe(lll) or Fe(ll) (Ahmann et al., 1997).
Field studies have demonstrated the importance of these
microorganisms in reducing  As(V) to As(lll) associated
with  As-contaminated  lake sediments  (Ahmann et al.,
1997); As(V) metabolism  under anaerobic conditions was
shown to be responsible for at least 14 percent of the car-
bon mineralization  in hypersaline Mono Lake, California,
where  arsenic concentrations up to 200  uM result from
hydrothermal inputs (Oremland  et al., 2000). Arsenic(V)
reducers have  been found in  freshwater,  alkaline, and
hypersaline environments but, thus far, only circumstantial
evidence points to their occurrence in acidic environments
(Oremland et al., 2001).

Colloidal Transport

Laboratory studies suggest that certain conditions within
the subsurface may develop that enhance arsenic transport
via mobile colloids.  Increased mobility of arsenic-bearing
colloidal material  may result either from changes in the
surface charge on colloids due to changes in subsurface
geochemistry (e.g., Puls and Powell, 1992;Tadanier et al.,
2005) or through deflocculation and resuspension of colloi-
dal material through dissolution of cementing agents within
the aquifer matrix (e.g., Ryan and Gschwend, 1990; Ryan
and Gschwend, 1992). Both processes would be facilitated
in aquifers impacted by organic contaminants where micro-
bial activity may be stimulated resulting in the generation of
reducing conditions and/or the production of low molecular
weight organic compounds that  partition  to fine-grained
sediments. Due to the affinity of arsenic to the surfaces of
iron oxyhydroxides that are commonly present as mineral
coatings on aquifer sediments, mobilization of these colloid-
sized solids could potentially contribute to arsenic mobility.
The distance of colloidal transport from an impacted zone
is uncertain, since colloid stability may change significantly
during transport from reduced to oxidized zones or due to
encounter with new sediment surfaces in unimpacted zones.
However, identification of this transport mechanism within
an arsenic plume may play an important role relative to the
management of the source zone or the establishment of a
monitoring system appropriate for determining the extent
of  colloidal transport.

Site Characterization

Determining the processes  controlling arsenic mobility in
ground water and forecasting the capacity and longevity for
attenuation is dependent  on understanding the chemical
processes controlling partitioning of arsenic onto aquifer
solids. The  aqueous and solid phase speciation of arse-
nic provides clues to  the processes controlling  solid-liquid
partitioning,  and, therefore,  changes  in mobility that may
accompany  chemical perturbations  within the  aquifer.
Table 6.1 indicates possible natural attenuation and mobi-
lization pathways for  arsenic.

Aqueous  Measurements

Determination of  arsenic chemical speciation in ground
water provides a basis for both assessing the factors con-
tributing to arsenic mobilization and the potential for arsenic
partitioning to aquifer solids within and down gradient from
the contaminant plume. The presence of arsenic as arsenite
or  thioarsenic species suggests that reductive processes
influence arsenic mobility.  In contrast, the predominance
of  the arsenate species in ground water suggests that de-
sorption processes due to elevated pH or competition from
other anions for sorption sites may be controlling arsenic
mobility. Significant concentrations of methylated or other
organo-arsenic species point to the potential influence
of  microbial  processes on arsenic mobility. Thus,  knowl-
edge of aqueous arsenic  speciation within the context of
site ground-water geochemistry within and outside of the
boundaries of the contaminant plume  is important relative
to  determining the current conditions controlling  arsenic
mobility (or immobility) as well as estimating the potential
for changes in arsenic mobility  that may coincide with
geochemical changes that occur as the aquifer returns to
pre-contamination conditions.
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Table 6.1  Natural attenuation pathways for arsenic.
 Attenuation Processes
     Mobilization Processes
           Characterization Approach
 Precipitation of metal
 arsenates or arsenites or
 precipitation of arsenic
 sulfides
Dissolution of metal arsenates/
arsenites due to change in pH;
dissolution of arsenic sulfides due
to increase in pH or shift from
reducing to oxidizing conditions.
Evaluation of arsenic speciation in ground water.
Determination of total As in the solid matrix and
suspected components in arsenic-bearing mineral.
Evaluation of mineral solubility relative to ground-
water chemistry and published solubility constants.
 Co-precipitation of ar-
 senic as a trace compo-
 nent in oxyhydroxides or
 sulfides of iron or man-
 ganese
Dissolution of host oxyhydroxide
due to decrease in pH or shift
from oxidizing to reducing condi-
tions; dissolution of host sulfide
due to shift from reducing to oxi-
dizing conditions.
Evaluation of arsenic speciation in ground water.
Evaluation of host precipitate formation relative to
existing ground-water chemistry; determination of
host mineral content in aquifer sediments via min-
eralogical characterization.  Evaluation of arsenic
solid phase partitioning using  sequential extraction
methodologies.
 Adsorption to iron oxyhy-
 droxides, iron sulfides, or
 other mineral surfaces
Desorption at high pH for oxyhy-
droxides and sulfides; complex-
ation/stabilization in the presence
of DOC.  Reductive dissolution of
iron hydroxides or oxidative dis-
solution of iron sulfides.
Evaluation of arsenic speciation in the aqueous
phase. Determination of total As in the solid ma-
trix.  Evaluation of arsenic solid phase partitioning
using sequential extraction methodologies.  Batch
and column testing to determine As uptake capac-
ity of site-specific aquifer  materials with variable
geochemical conditions.
The total concentration of arsenic in aqueous samples can
be assessed by a number of methods depending on the re-
quired sensitivity to meet data quality objectives (Melamed,
2004; Melamed,  2005).  Colorimetric detection based on
the molybdenum blue method has been employed to mea-
sure inorganic arsenic for concentrations of approximately
2 ug L1  and higher (Johnson, 1971; Woolson et al., 1971;
Johnson and Pilson, 1972). Determination of total arsenic
may also be achieved using silver dithyldithiocarbamate
as a colorimetric reagent following conversion of arsenic
to arsine gas (Aggett and Aspell, 1976).  However, these
methods typically require conversion of all  arsenic species
in a sample to a  single aqueous or gaseous species prior
to colorimetric determination.  These  methods have been
applied  in the field and commercial kits are available (e.g.,
Steinmaus  et al., 2006). The colorimetric methods suffer
from analytical interferences (Stauffer, 1983; Frenzel et al.,
1994) and  in some cases require the use of hazardous
substances for test application.

More commonly, total arsenic is  determined  by  element-
specific atomic absorption/emission or mass spectrometry
(Eaton et al.,  1998).  Samples can be introduced either
directly  into the  spectrometer or following conversion of
arsenic  to a hydride gas. These methods are less prone
to analytical interferences, but complex sample  matrices
can  interfere with the atomization or hydride conversion
steps.   In addition, both atomic  and  mass spectrometry
may suffer  from  spectral interferences either from other
elements within the sample or from molecular species that
absorb at similar energy or possess the mass employed
for quantification (Feldmann et al., 1999).
                             There  are generally two  approaches  to  analytical  de-
                             termination of aqueous arsenic speciation:  1) chemical
                             separation of individual species in the field and subse-
                             quent determination of total arsenic in each fraction and
                             2) preservation of the in-situ arsenic speciation followed by
                             chemical separation and quantification of each specie in
                             the laboratory. Numerous studies suggest that field sepa-
                             ration of the various arsenic species can be achieved for
                             drinking water supplies via the use of exchange resins that
                             selectively pass or retain specific species (e.g., Wilkie and
                             Hering, 1998; Le et al., 2000; Yalcin and Le, 2001). While
                             this method is attractive, there  are limitations in species
                             separation (Miller et al.,  2000) and potential interferences
                             from competing cations or anions in contaminated water.
                             This latter factor has not been sufficiently addressed for the
                             complex chemistry frequently encountered in contaminant
                             plumes. Thus, it is recommended that method validation
                             be performed on a site-specific  basis for all ground-water
                             compositions that may be encountered.

                             The use of chemical and/or physical methods to preserve
                             in-situ arsenic speciation has been investigated for a range
                             of sample holding times (e.g., Haswelletal., 1985; Palacios
                             etal., 1997; Eaton etal., 1998; Hall et al., 1999; Gallagher
                             et al., 2004).  One recurring observation from all studies is
                             the importance of steps taken to prevent the precipitation of
                             hydrous ferric oxide in samples collected from Fe-rich anoxic
                             water in order to prevent the loss  of arsenic from solution. A
                             recent study provides a comprehensive review of previously
                             published research along with the results of a contempo-
                             rary study conducted using water samples from numerous
                             sources (McCleskey et al., 2004). The analysis presented
                                                     62

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in this study reconciles many of the inconsistencies from
historical studies that examined the use of various chemi-
cal reagents for preservation of  arsenic species in water.
The authors concluded that acidification and storage of the
sample at 4 C in a container that excludes light effectively
preserves the As(lll)/As(V) ratio for adequate periods for a
wide range of chemical compositions of non-sulfidic ground
water. In sulfidic waters, it appears that most of the methods
for preservation of both total and individual arsenic species
are inadequate either due to  susceptibility of thioarsenic
species to oxygen exposure or the precipitation of arse-
nic sulfides upon acidification (Smieja and Wilkin,  2003).
However, preservation approaches have been developed
for sulfidic waters that  support  reliable measurement  of
total arsenic (Smieja and Wilkin, 2003; Samanta and Clif-
ford, 2006). Samanta and Clifford (2006) also demonstrate
that anion exchange may be  used for field separation  of
arsenite and thioarsenic species in reduced water when
these are the dominant arsenic species. Regardless of the
method of preservation employed,  it is recommended that
performance validation for site-specific water chemistry be
undertaken through use of ground-water samples  spiked
with  known standards in the field.  The use of controlled
laboratory solutions provides an inadequate test for the
complexity of water chemistries that are typically encoun-
tered within contaminated ground-water systems.

Arsenic speciation in preserved aqueous samples  can
be quantified following chemical  separation or masking  of
the various species. The molybedenum blue colorimetric
method has been employed on natural samples to quantify
As(lll) and  As(V) following manipulation of the oxidation
state of  As(lll) (Johnson and Pilson, 1972).  This method
may be employed in the field, but is subject to interferences
from the complex chemical matrices commonly encountered
in contaminated ground water. More commonly,  aqueous
speciation is achieved through chromatographic separation
of the various species followed by determination of arsenic
by atomic absorption/fluorescence or mass spectrometry
(Gabon and Gabon, 2000; Gomez-Ariza et al.,  2000).

Solid Phase Measurements

The  implementation of  an analytical approach to identify
arsenic speciation in a soil sample is a challenging process.
The accuracy of the analytical finding  is dependent on the
method  of  sample collection/preservation and the tools
used to identify the mechanism of arsenic partitioning.  It
is recommended that the analytical protocol be designed
to address  the redox sensitivity of arsenic and the solid
phase(s) to which arsenic may  be partitioned.  Tools  to
evaluate the mechanism of arsenic solid phase partitioning
range in complexity from relatively simple chemical extrac-
tions to advanced spectroscopic techniques.

Bulk solid phase arsenic concentration can be determined
directly on the solid sample by X-ray fluorescence spectrom-
etry, neutron activation  analysis, or following chemical di-
gestion and analysis of arsenic in the resultant liquid  phase.
Neutron activation analysis  is not commonly employed
due to the scarcity of neutron sources required to irradiate
the sample.  X-ray fluorescence is the most attractive ap-
proach due to the relative ease of sample preparation, but
there are potential spectral interferences. For example, the
presence of elevated concentrations of lead can interfere
with arsenic quantification due to overlap of the As Ka and
the Pb La fluorescence peaks (Wegrzynek and Holynska,
1993). There are approaches to correct for this interference
provided the Pb  L(3 fluorescence peak is measurable and
free from interference. In addition, it is recommended that
sample preparation via fusion with lithium metaborate be
avoided due to the possible loss of arsenic through vola-
tilization at the fusion temperature (Alvarez, 1990). There
are various chemical digestion methods for determination of
total arsenic (Kane, 1995; Chen and Ma, 1998), but these
procedures do not always provide complete recovery.  In
addition, arsenic is a  relatively volatile element  (Yang  et
al., 1998),  so  sample heating to improve recovery can
lead to  losses  without the appropriate precautions. Both
approaches to  quantifying total solid phase arsenic can be
reliably applied provided the analyst is aware of the pos-
sible analytical artifacts.

There have been many applications of sequential extraction
schemes to assess the speciation of  solid phase arsenic
(e.g., Moore et al., 1988; McLaren et al.,  1998; La Force
et al., 2000; Lumsdon et al.,  2001).  This approach has
been employed primarily due to the inability to apply con-
ventional laboratory techniques such as X-ray diffraction
or infrared spectroscopy for detection of trace constituents
within a soil/sediment and the limited availability of more
advanced element-specific spectroscopies such  as X-ray
absorption spectroscopy at a synchrotron radiation facility.
However, the lack of validated procedures and reference
materials to test the ability of the various chemical extract-
ants that have been employed limits the validity of selective
extraction schemes. There are documented instances that
clearly show selective extractants are incapable of targeting
arsenic associated with specific solid components within a
soil/sediment (e.g., Gruebel et al., 1988; Wilkin and Ford,
2002). There is  no single extraction protocol that can be
recommended to chemically speciate solid phase arsenic.
It is recommended that protocols be developed on a site-
specific basis  with knowledge of the prevailing aqueous
geochemistry and mineralogy of the ground-water aquifer
(e.g., Ford et al., 2005).

Long-Term  Stability and Capacity

The long-term  stability of arsenic immobilized onto aquifer
solids will depend on the prevailing ground-water chemis-
try over time relative to the  conditions that existed at the
time of immobilization.  If there are  significant  changes
in ground-water  chemistry following  immobilization, then
the potential exists for remobilization of arsenic with the
establishment of a new mobile plume. A general example
of this type of  situation would be if arsenic was originally
immobilized through partitioning to mineral sulfides within
aquifer sediments under sulfate-reducing  conditions. Ar-
senic associated with sulfide minerals may be re-mobilized
due to sulfide oxidation if the aquifer were to return to more
oxidizing conditions. There  are several examples of field
                                                     63

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studies that document the potential for release of arsenic
bound to aquifer solids subsequent to changes in ground-
water chemistry (Table 6.2).  The examples provided in
Table 6.2 provide general perspectives of factors that may
increase arsenic mobility relative to the original mode for
immobilization.  While these examples are not exhaustive
of all  possible situations,  they do point to the importance
of establishing a monitoring program that is designed to
consider the potential impact of changes in site geochem-
istry relative to the conditions under which arsenic was
attenuated within the aquifer. Ultimately,  the evolution of
biogeochemical conditions in the subsurface will dictate the
success of the MNA remedy.

Quantifying the attenuation capacity will also necessitate
an understanding of the  specific attenuation pathway(s).
Attenuation capacity, for example, could be related to the
extent that pH is buffered, the availability of sorptive sites in
aquifer materials, or to the supply of electron donors needed
to sustain microbially mediated redox conditions. For any
proposed attenuation mechanism, there will be assumptions
built into capacity estimations, so that uncertainty analysis
is recommended to support capacity calculations.

Tiered Analysis

Determination of  the viability of arsenic remediation in
ground water via monitored natural attenuation will depend
upon  proper  assessment of contaminant loading to the
aquifer and prevailing geochemistry and mineralogy within
the  contaminant plume and the  down gradient  zone prior
to the point(s) of compliance. The goal of site assessment
is to demonstrate the process(es) controlling arsenic se-
questration onto aquifer solids and the  long-term stability
                     of solid phase arsenic as a function of existing and antici-
                     pated ground-water chemistry. A recent technical review
                     highlights limited instances where natural attenuation was
                     chosen as a component of the remedy for cleanup of arsenic
                     in ground water (Reisinger et al., 2005). However, as noted
                     in the  review,  site characterization conducted to support
                     selection of natural attenuation as a remedy at these sites
                     did not fully identify the site-specific immobilization process
                     and the long-term stability of the sequestered arsenic. The
                     following tiered analysis structure for site characterization
                     provides a technically defensible approach to evaluate
                     candidate sites and define the potential limitations of MNA
                     as part of a remedy for ground-water cleanup.

                     Tier I-  Site characterization under Tier I will involve demon-
                     stration that the plume is static or shrinking, has not reached
                     compliance boundaries, and does not impact existing water
                     supplies.  Once this  is established through ground-water
                     characterization, evidence is collected to demonstrate As
                     partitioning to aquifer solids within the plume. If natural
                     attenuation processes are active throughout the plume, it
                     is anticipated that this would be reflected in an observed
                     increase  in solid  phase  concentrations within  regions
                     of the  plume  with higher aqueous concentrations, e.g.,
                     near the source term. This field partitioning data may be
                     supplemented by geochemical modeling that incorporates
                     measured water chemistry (e.g.,  pH,  Eh,  and major ion
                     chemistry) throughout the plume to assess the potential for
                     solubility control by metal arsenates/arsenites or arsenic
                     sulfides.  Since identification of the chemical speciation of
                     arsenic in water samples and aquifer sediments is critical
                     towards determining the attenuation mechanism(s), it is
                     recommended that precautions be taken to preserve arsenic
Table 6.2  Examples of arsenic mobilization due to shifts in ground-water chemistry.
      Location and Setting
      Original Conditions
      (Arsenic Immobile)
           Altered Conditions
           (Arsenic Mobilized)
 Unconsolidated aquifer consisting
 of glacial outwash, glaciomarine
 clay, and till overlying bedrock;
 New Hampshire a
Arsenic bound to poorly crystal-
line iron hydroxides in glacioma-
rine clay under oxidizing condi-
tions (Eh ~ 400 mV)
Reductive dissolution of iron hydroxides due
to microbially driven Fe-reducing condi-
tions stimulated by organic carbon in landfill
leachate
 Suwannee Limestone, Upper
 Floridan consolidated aquifer;
 Florida b
Arsenic hosted in pyrite that is
most abundant in high porosity
zones; anoxic aquifer
Oxidative dissolution of pyrite due to injec-
tion of oxygenated water
 Unconsolidated, glacial out-
 wash aquifer consisting of
 coarse(quartz) sand and gravel
 with  Fe and Al oxide and/or
 silicate mineral coatings; New
 Hampshire c
Arsenic adsorbed to mineral
coatings under oxidizing condi-
tions
Adsorbed arsenic mobilized due to desorp-
tion in the presence of elevated phosphate
derived from sewage disposal via land ap-
plication coupled with reductive dissolution
of Fe-bearing mineral coatings
1 Delemeos et al., 2006;b Price and Pichler, 2006;  Kent and Fox, 2004
                                                     64

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speciation during collection, preservation, and processing of
collected samples (See recommendations and application
of these methods in Wilkin et al., 2002, Ford et al., 2005,
and Wilkin, 2006).  Identification of active sequestration to
prevent arsenic migration in ground-water provides justifica-
tion for proceeding to Tier II characterization efforts.

Tier II - Under Tier II, the apparent rate and mechanism(s)
of attenuation are determined. Estimates of a site  attenu-
ation rate(s) can be assessed via a well transect along
the ground-water flow path.  In addition, time-series data
may be collected at one or more monitoring points within
the plume. This information  will allow assessment of the
relative timescales for contaminant immobilization and fluid
transport and determination of whether remediation objec-
tives can be met within the required regulatory time frame.
In addition, the mechanism(s) for attenuation need to be
identified  under  this stage of site characterization.  This
effort may require determination of the chemical speciation
of aqueous and solid phase As, which may be approached
according to the following scheme:

    1)  Determination of arsenic solution  speciation
       via direct analytical measurements in combi-
       nation with speciation calculations based  on
       characterized  ground-water chemistry;
    2)  Calculation of the saturation state of ground
       water  relative to measured aqueous chem-
       istry complimented by the possible  isolation
       of  discrete As  mineral  phases via density
       separations (or other schemes) in regions of
       the aquifer with highest solid phase concentra-
       tions;
    3)
    4)
        Determination of aquifer mineralogy to deter-
        mine the relative abundance of components
        with documented capacity for As sorption (e.g.,
        Amonette, 2002);
        Identification of arsenic association(s) with
        the various solid phase components of aquifer
        solids through combination of chemical extrac-
        tions with microscopic/spectroscopic confirma-
        tion of phase associations,  and;
    5)   Demonstration of concurrence between the site
        conceptual model and mathematical model(s)
        that describe arsenic removal mechanism(s).
It is recommended that  identification of arsenic chemical
speciation in aqueous and solid matrices be conducted us-
ing samples collected in a manner that preserves the in-situ
mineralogy and speciation of arsenic. The demonstration
of concurrence  between  conceptual  and mathematical
models describing arsenic transport  will entail development
of site-specific parameterization of the chemical processes
controlling arsenic solid  phase partitioning.

Tier III - Once the partitioning mechanism(s) have been
identified for the site, the subsequent characterization effort
under Tier  III will involve determination of the  stability of
immobilized As and the  capacity of the aquifer to sustain
continued uptake.  It is recommended that the stability of
immobilized As be tested based on the anticipated evolution
of ground-water chemistry concurrent with plume shrink-
age. For example, changes in ground-water redox potential
due to oxygen  intrusion or the occurrence of anaerobic
microbial processes can exert a significant influence on As
partitioning to iron sulfides or iron oxyhydroxides, respec-
tively. Therefore,  it  is recommended that sediment leach
tests be conducted to characterize the magnitude of As
mobilization as a function of ground-water redox chemistry
representative of  existing and anticipated site conditions.
This may involve the use of microcosm tests that stimulate
in-situ microbial populations toward the development of
redox conditions considered deleterious for continued As
immobilization.  It is recommended that the capacity for
As uptake onto aquifer solids be determined relative to the
specific mechanism(s) identified in Tier II. For example, if
site characterization under Tier II indicated that co-precipita-
tion of As with iron sulfide due to microbial degradation of
organic  compounds coupled with sulfate reduction occurs
within the aquifer, then it is recommended that the mass
distribution of organic carbon,  sulfate and ferrous iron to
support this reaction within the aquifer be determined. This
site-specific capacity can then be compared to As mass
loading  within the plume in order to assess the longevity
of the natural  attenuation process.  If site-specific tests
demonstrate the stability of immobilized As and sufficient
capacity within the aquifer to sustain As attenuation, then
the site  characterization effort can progress to Tier IV.  For
cases where contaminant stability is sufficient but aquifer
capacity is insufficient for capture of the entire plume, then
a determination of the benefits of contaminant source re-
duction  may be necessary.

Tier IV - Finally, under Tier IV a monitoring plan is estab-
lished along with  contingency plans in the event of MNA
failure.  It is recommended that the monitoring plan be
designed to establish both continued plume stability and to
identify  changes in ground-water chemistry that may lead
to re-mobilization  of attenuated As. The specific chemical
parameters to be monitored will include those  identified
under Tier III that  may halt As partitioning to  aquifer sedi-
ments and/or result in solubilization of either discrete As
precipitates or  aquifer minerals  that sequester As from
ground  water.   For  example, solution phase parameters
that could alter either As precipitation or adsorption include
increases  in the concentration  of competing anions, such
as phosphate, in combination with changes in ground-wa-
ter pH.  In contrast, the concentration of dissolved iron or
sulfate may indicate the dissolution of an important sorptive
phase within the aquifer (e.g., reductive dissolution of  iron
oxyhydroxides or oxidative dissolution of sulfides). Changes
in these parameters may occur prior to observed changes
in solution As and,  thus, serve as monitoring triggers for
potential MNA failure.  In this instance, a contingency plan
can be  implemented that incorporates interventive strate-
gies to arrest possible plume expansion beyond compliance
boundaries. Possible strategies to prevent plume expansion
include  pump and treat operations, installation of reactive
barriers to enhance uptake capacity perpendicular to the
direction of plume advance, or enhancement of natural at-
tenuation processes within the aquifer through the injection
of soluble reactive components.
                                                     65

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References

Aggett, J. and A.C. Aspell. Determination of arsenic (III)
    and total arsenic by the silver diethyldithiocarbamate
    method. Analyst 101: 912-913 (1976).
Ahmann, D., A.L. Roberts, L.R. Krumholz, and F.M.M. Mo-
    rel.  Microbe grows by reducing arsenic. Nature 371:
    750(1994).
Ahmann, D., H.F. Krumholz, H.F. Hemond, D.R. Lovley, and
    F.M.M.  Morel. Microbial mobilization of arsenic from
    sediments of the Aberjona watershed: Environmental
    Science and Technology 31: 2923-2930 (1997).
Alvarez, M. Glass disk fusion method for the X-ray fluo-
    rescence  analysis of rocks  and silicates. X-ray Spec-
    trometry 19: 203-206 (1990).
Amonette,  J.E.   Methods for determination  of mineral-
    ogy and environmental availability. In So/7 Mineralogy
    with Environmental Applications. J.B. Dixon and D.G.
    Schulze (Eds.), Soil Science Society of America Book
    Series.  Soil Science Society of America, Inc., Madison,
    Wl, pp.  153-197(2002).
Anderson, L.C. and K.W. Bruland. Biochemistry of arsenic
    in natural waters: the importance of methylated species.
    Environmental Science and Technology 25: 420-427
    (1991).
Bostick, B.C., S. Fendorf, and G.E. Brown Jr. In situ analysis
    of thioarsenite complexes in neutral to alkaline arsenic
    sulphide solutions. Mineralogical Magazine 69: 781-
    795 (2005).
Gabon, J.Y. and N. Gabon. Determination of arsenic spe-
    cies in  seawater  by flow injection hydride generation
    in situ collection followed by graphite furnace  atomic
    absorption spectrometry: Stability of As(lll). Analytica
    Chimica Acta 418: 19-31 (2000).
Chen, M. and  L.Q. Ma. Comparison of four USEPA diges-
    tion methods for trace metal analysis using certified
    and Florida soils. Journal of Environmental Quality
    27: 1294-1300(1998).
Cheng, Z.,  A. van Geen,  R. Louis, N. Nikolaidis, and  R.
    Bailey. Removal of methylated arsenic in groundwater
    with iron filings. Environmental Science and Technology
    39: 7662-7666 (2005).
Chiu, V.Q. and J.G. Hering. Arsenic adsorption and oxida-
    tion at  manganite surfaces. 1. Method for simultane-
    ous determination of adsorbed and dissolved arsenic
    species. Environmental Science and Technology 34:
    2029-2034 (2000).
Clarke, M.B. and G.R. Helz. Metal-thiometalate transport
    of biologically active trace elements in sulfidic environ-
    ments. 1. Experimental evidence for copper thioarsenite
    complexing. Environmental Science and  Technology
    34: 1477-1482 (2000).
Cullen, W.R. and K.J. Reimer.  Arsenic speciation  in the
    environment. Chemical Reviews 89: 713-764 (1989).
de Vitre, R., N. Belzile, and A. Tessier. Speciation and ad-
    sorption of arsenic on diagenetic iron  oxyhydroxides.
    Limnology and Oceanography 36: 1480-1485 (1991).
Davis, J.A., and D.B. Kent. Surface complexation modeling
    in aqueous geochemistry. In Mineral-Water Interface
    Geochemistry.  M.F Hochella and A.F White (Eds.),
    Washington DC. Mineralogical Society of America, pp.
    177-260(1990).
Delemeos, J.L., B.C. Bostick, C.E. Renshaw, S. Sturup, and
    X. Feng. Landfill-stimulated iron reduction and arsenic
    release at the Coakley Superfund Site (NH). Environ-
    mental Science and Technology 40: 67-73 (2006).
Dzombak, D.A. and F.M.M. Morel.  Surface Complexation
    Modeling: Hydrous Ferric Oxide. John Wiley and Sons,
    New York, NY, 393 pp. (1990).
Eaton, A., H.C.Wang, and J.Northington.  Analytical Chem-
    istry of Arsenic in Drinking Water. AWWA Research
    Foundation and American Water Works Association,
    Denver, CO, 148 pp. (1998).
Feldmann, I., N. Jakubowski, C.Thomas, and D. Stuewer.
    Application of a hexapole collision and reaction cell
    in ICP-MS. Part II: Analytical figures  of merit and first
    applications. Fresenius Journal of Analytical Chemistry
    365:422-428(1999).
Ferguson, J.F. and J. Gavis. A review of the arsenic cycle in
    natural waters.  Water Research 6: 1259-1274 (1972).
Ford, R.G. Rates of hydrous ferric oxide crystallization and
    the influence on coprecipitated arsenate. Environmental
    Science and Technology 36: 2459-2463 (2002).
Ford, R.G.   The  Impact of Ground Water-Surface Water
    Interactions on  Contaminant Transport with Application
    to an Arsenic Contaminated Site. EPA Environmental
    Research Brief, EPA/600/S-05/002.   Cincinnati, OH:
    U.S. Environmental  Protection Agency,  (http://www.
    epa. gov/ada/down load/brief s/epa_600_s05_002. pdf)
    (2005).
Ford, R.G., R.T. Wilkin, and G. Hernandez. Arsenic cycling
    within the water column of a small lake receiving con-
    taminated ground-water discharge. Chemical Geology
    228: 137-155(2006).
Ford, R.G., R.T.Wilkin, K.G.Scheckel, C.J.Paul, FBeck, P.
    Clark, and  T  Lee.  Field Study of the Fate of Arsenic,
    Lead,  and Zinc at the Ground-Water/Surface-Water
    Interface. EPA Report, EPA/600/R-05/161. Cincinnati,
    OH: U.S. Environmental Protection Agency, (http://
    www.epa.gov/ada/download/reports/600R05161/
    600R05161.pdf) (2005).
Foster, A.L., G.E. Brown Jr., T.N. Tingle, and G.A. Parks.
    Quantitative arsenic speciation in mine tailings  using
    X-ray absorption spectroscopy. American Mineralogist
    83:553-568(1998).
Frenzel, W., F  Titzenthaler, and S. Elbel. Selective deter-
    mination of arsenite by flow injection spectrophotometry.
    Talanta 41: 1965-1971 (1994).
Gallagher, PA., C.A. Schwegel, A. Parks, B.M. Gamble, L.
    Wymer, and J.T Creed.  Preservation of As(lll) and
    As(V) in drinking water supply samples from across the
    United States using EDTA and acetic acid as a means
    of minimizing iron-arsenic coprecipitation. Environmen-
    tal Science and Technology 38: 2919-2927 (2004).
                                                    66

-------
Goldberg, S. and R.A. Glaubig. Anion sorption on a calcare-
    ous, montmorillonitic soil - Arsenic. So/7 Science Society
    of America Journal 52: 1297-1300 (1988).
Gomez-Ariza, J.L., D. Sanchez-Rodas, I. Giraldez, and E.
    Morales. A comparison between ICP-MS and AFS de-
    tection for arsenic speciation in environmental samples.
    Talanta 51: 257-268 (2000).
Gruebel, K.A., J.A. Davis, and J.O. Leckie. The feasibility of
    using sequential extraction techniques for arsenic and
    selenium  in soils and sediments.  So/7 Science Society
    of America Journal 52: 390-397 (1988).
Hall, G.E.M.,  J.C. Pelchat,  and G. Gauthier. Stability of
    inorganic  arsenic(lll) and arsenic(V) in water samples.
    Journal of Analytical Atomic Spectrometry 14:205-213
    (1999).
Harrington, J. M., S.E. Fendorf, and R.F. Rosenzweig. Biotic
    generation of  arsenic(lll) in  metal(loid)-contaminated
    freshwater lake sediments. Environmental Science and
    Technology 32: 2425-2430 (1998).
Haswell,  S.J., P.  O'Neill, and  K.C.C.  Bancroft. Arsenic
    speciation in  soil-pore waters from mineralized and
    unmineralized areas of south-west England. Talanta
    32:69-72 (1985).
Hoeft, S.E., F Lucas, J.T. Hollibaugh, and R.S. Oremland.
    Characterization of microbial  arsenate  reduction in
    the anoxic bottom  waters of Mono Lake, California.
    Geomicrobiology Journal 19: 23-40 (2002).
Hollibaugh, J.T., S. Crini, H. Gurleyuk, R.  Jellison, S.B. Joye,
    G. LeCleir, C.  Meile, L. Vasquez, and D. Wallschlager.
    Arsenic speciation in Mono Lake, California: Response
    to seasonal stratification and anoxia. Geochimica et
    Cosmochimica Acta 69: 1925-1937 (2005).
Hounslow, A.W. Ground-water geochemistry:  arsenic in
    landfills. Ground Water 18: 331-333 (1980).
Jackson, B.P  and P.M.  Bertsch.  Determination of arsenic
    speciation in poultry wastes by IC-ICP-MS. Environmen-
    tal Science and Technology 35: 4868-4873 (2001).
Jackson, B.P.  and W.P  Miller. Effectiveness of  phosphate
    and hydroxide for desorption of arsenic and selenium
    species from  iron  oxides.  So/7  Science  Society of
    America Journal64: 1616-1622 (2000).
Johnson, D.L. Simultaneous determination of arsenate and
    phosphate in  natural waters. Environmental Science
    and Technology 5: 411 -414 (1971).
Johnson, D.L. and M.E.  Pilson. Spectrophotometric deter-
    mination of arsenite, arsenate, and phosphate in natural
    waters. Analytica Chimica Acta 58: 289-299 (1972).
Jones, C.A., H.W.  Langner, K. Anderson, T.R. McDermott,
    and W.P. Inskeep. Rates of microbially mediated arse-
    nate  reduction and solubilization. So/7 Science Society
    of America Journal 64: 600-608 (2000).
Kane, J.S. Leach data vs. total: which is relevant for SRMs?
    Fresenius Journal of Analytical Chemistry 352: 209-
    218(1995).
Kent,  D.B. and P.M.  Fox. The  influence of  groundwater
    chemistry on arsenic concentrations and speciation in
    a quartz sand and gravel aquifer. Geochemical Trans-
    actions 5: 1-12 (2004).
Korte, N.E. and Q. Fernando. A review of arsenic (III) in
    groundwater. Critical Reviews in Environmental Control
    21: 1-39 (1991).
La Force, M.J., C.M. Hansel, and S. Fendord. Arsenic spe-
    ciation, seasonal  transformations, and co-distribution
    with iron in a mine waste-influenced palustrine emer-
    gent wetland. Environmental Science and Technology
    34: 3937-3943 (2000).
Lackovic, J., N. Nikolaidis, and G. Dobbs. Inorganic arsenic
    removal by zero-valent iron. Environmental Engineering
    and Science 17: 29-39 (2000).
Le, X.C., S.Yalcin, and M. Ma. Speciation of submicrogram
    per liter levels of arsenic in water: On-site species sepa-
    ration integrated with sample collection. Environmental
    Science and Technology 34: 2342-2347 (2000).
Lowers, H.A., G.N. Breit, A.L. Foster, J. Whitney, J. Yount,
    Md.N. Uddin, and Ad.A. Muneem. Arsenic incorpora-
    tion into authigenic pyrite, Bengal  Basin sediment,
    Bangladesh. Geochimica et Cosmochimica Acta 71:
    2699-2717(2007).
Lumsdon, D.G., J.C.L.  Meeussen, E. Paterson, L.M. Garden,
    and P. Anderson.  Use of solid  phase characterisation
    and chemical modeling for assessing the behaviour of
    arsenic in  contaminated  soils. Applied Geochemistry
    16:571-581 (2001).
Maeda, S. Biotransformation of arsenic in the freshwater en-
    vironment.  In Arsenic in the Environment. Part I: Cycling
    and Characterization. J.O. Nriagu (Ed.), John Wiley &
    Sons, Inc., New York, NY, pp. 155-187 (1994).
Mandal, B.K. and K.T Suzuki. Arsenic round the world: a
    review. Talanta 58: 201-235 (2002)
Manning, B.A., S.E. Fendorf, and  S. Goldberg.  Surface
    structures  and stability  of arsenic(lll)  on goethite:
    Spectroscopic  evidence  for inner-sphere complexes.
    Environmental Science and Technology 32:2383-2388
    (1998).
McCleskey, R.B., O.K. Nordstrom, and A.S. Maest. Pres-
    ervation of water samples for arsenic(lll/V) determina-
    tions: an evaluation of the literature and new analytical
    results. Applied Geochemistry 19: 995-1009 (2004).
McGeehan, S.L., S.E. Fendorf, and D.V. Naylor. Alteration
    of arsenic sorption in flooded-dried soils. So/7 Science
    Society of America Journal 62: 828-833 (1998).
McLaren, R.G., R. Naidu, J. Smith, and K.G.Tiller. Fraction-
    ation and distribution of arsenic in soils contaminated
    by cattle dip.  Journal of Environmental Quality 27:
    348-354(1998).
McNeill,  L.S. and  M. Edwards. Arsenic removal during
    precipitative softening. Journal of Environmental En-
    gineering 123(5):  453-460 (1997).
                                                    67

-------
Melamed, D.  Monitoring Arsenic in the Environment: A
    Review of Science and Technologies for Field Measure-
    ments and Sensors, EPA/542/R-04/002  . Washington,
    D.C.: U. S. Environmental Protection Agency, (http://
    cluin.org/download/char/arsenic_paper.pdf) (2004).
Melamed, D.   Monitoring arsenic in the environment: a
    review of science and technologies with the potential
    for field measurements. Analytica Chimica Acta 532:
    1-13(2005).
Meng, X., S. Bang, and G.P. Korfiatis.  Effects of silicate,
    sulfate, and carbonate on arsenic removal by ferric
    chloride. Water Research 34: 1255-1261 (2000).
Miller, G.P, D.I. Norman, and PL. Frisch. A comment on
    arsenic species separation using ion exchange. Water
    Research 34(4): 1397-1400 (2000).
Moore, J.N. Contaminant mobilization resulting from redox
    pumping in a metal-contaminated river-reservoir sys-
    tem. In Environmental Chemistry of Lakes and Reser-
    voirs. L.A. Baker (Ed.),  American Chemical  Society,
    Washington DC, pp. 451-471 (1994).
Moore, J.N., W.H. Ficklin, and C. Johns.  Partitioning of
    arsenic and  metals  in  reducing  sulfidic  sediments.
    Environmental Science  and Technology 22: 432-437
    (1988).
Morse, J.W. Interactions of  trace metals with authigenic
    sulfide minerals: implications for their  bioavailability.
    Marine Chemistry 46: 1-6 (1994).
Nordstrom, O.K. and D.G. Archer. Arsenic thermodynamic
    data and environmental geochemistry.  In Arsenic in
    Ground Water: Geochemistry and Occurrence. A.M.
    Welch and K.G. Stollenwerk (Eds.), Kluwer Academic
    Publishers, Boston, MA, pp. 1-26 (2003).
O'Day, PA., D. Vlassopoulos, R. Root, and N.  Rivera. The
    influence of sulfur and iron on dissolved arsenic con-
    centrations in the shallow subsurface under changing
    redox conditions. Proceedings of the National Academy
    of Sciences 101:13703-13708 (2004).
Oremland, R.S.,  PR. Dowdle,  S.E.   Hoeft, J.O. Sharp,
    J.K. Schaefer, L.G. Miller, J.S. Blum, R.L.  Smith, N.S.
    Bloom, and D.  Wallschlaeger. Bacterial dissimilatory
    reduction of arsenate and sulfate in meromictic Mono
    Lake, California.  Geochimica  et Cosmochimica Acta
    64: 3073-3084 (2000).
Oremland, R.S., S.E. Hoeft, J.M. Santini, N. Bano,  R.A.
    Hollibaugh, and J.T.  Hollibaugh. Anaerobic oxidation
    of arsenite in Mono Lake water and by a facultative,
    arsenite-oxidizing chemoautotroph. Applied and Envi-
    ronmental Microbiology 68: 4795-4802 (2002).
Oremland, R.S., O.K. Newman, B.W. Kail, and J.F. Stolz.
    Bacterial respiration  of arsenate and its significance
    in the environment.  In Environmental Chemistry of
    Arsenic. W.T. Frankengerger,  (Ed.),  Marcel  Dekker,
    New York, pp. 273-295 (2001).
Palacios, M.A., M. Gomez, C. Camara,  and M.A. Lopez.
    Stability  studies  of arsenate,  monomethylarsonate,
    dimethylarsinate, arsenobetaine and arsenocholine in
    deionized water, urine and clean-up dry residue from
    urine samples  and determination by liquid chroma-
    tography  with  microwave-assisted oxidation-hydride
    generation atomic absorption spectrometric detection.
    Analytica Chimica Acta 340: 209-220 (1997).
Peryea, FJ. Phosphate-induced release of arsenic from
    soils contaminated  with lead arsenate. So/7 Science
    Society of America Journal 55: 1301-1306 (1991).
Polizzotto, M.L., C.F Harvey, S.R. Sutton, and S. Fendorf.
    Processes conducive to the release and transport of
    arsenic into aquifers of Bangladesh. Proceedings of
    the National Academy of Sciences 102:18819-18823
    (2005).
Polizzotto, M.L., C.F. Harvey, G. Li, B. Badruzzman,  A. AN,
    M. Newville, S.  Sutton,  and S. Fendorf. Solid-phases
    and desorption processes of arsenic within Bangladesh
    sediments. Chemical Geology 228: 97-111 (2006).
Price, R.E. and T. Pichler. Abundance and  mineralogical
    association of  arsenic in the Suwannee Limestone
    (Florida): Implications for arsenic release during wa-
    ter-rock interaction. Chemical  Geology 228:  44-56
    (2006).
Puls,  R.W.  and R.M. Powell. Transport of inorganic col-
    loids through natural aquifer material: Implications for
    contaminant  transport.  Environmental  Science and
    Technology 26: 614-621 (1992).
Redman, A.D., D.L. Macalady,  and  D. Ahmann.  Natural
    organic matter affects arsenic speciation and sorption
    onto hematite. Environmental Science and Technology
    36: 2889-2896 (2002).
Reisinger, H.J., D.R. Burris, and J.G. Hering. Remediating
    subsurface arsenic contamination with monitored natu-
    ral attenuation. Environmental Science and Technology
    39: 458A-464A (2005).
Ryan, J.N. and P.M. Gschwend.  Colloid mobilization  in two
    Atlantic Coastal Plain aquifers: Field studies.  Water
    Resources Research 26: 307-322 (1990).
Ryan, J.N. and P.M. Gschwend.  Effect of iron diagenesis
    on the transport of colloidal clay in an unconfined sand
    aquifer. Geochimica et Cosmochimica Acta 56:  1507-
    1521 (1992).
Samanta, G. and D.A. Clifford. Influence of sulfide (S2") on
    preservation  and speciation  of inorganic arsenic in
    drinking water. Chemosphere 65: 847-853 (2006).
Santini, J.M., L.I. Sly, R.D. Schnagl, and J.M. Macy. A new
    chemolithoautotrophic  arsenite-oxidizing bacterium
    isolated from a gold mine: phylogenetic,  physiologi-
    cal, and preliminary biochemical studies. Applied and
    Environmental Microbiology 66: 92-97 (2000).
Senn, D.B. and H.F Hemond. Nitrate controls on iron and
    arsenic in an urban lake.  Science 296:  2373-2375
    (2002).
Smedley, PL. and D.G. Kinniburgh. A review of the source,
    behaviour and distribution of arsenic in natural waters.
    Applied Geochemistry 17: 517-568 (2002).
                                                    68

-------
Smieja, J.A. and R.T. Wilkin. Preservation of sulfidic waters
    containing dissolved As(lll). Journal of Environmental
    Monitoring 6: 913-916 (2003).
Stauffer, R.E. Determination of arsenic and phosphorus
    compounds in groundwaterwith reduced molybdenum
    blue. Analytical Chemistry 55: 1205-1210 (1983).
Steinmaus, C.M., C.M. George, D.A. Kalman, and A.M.
    Smith. Evaluation of two new arsenic field test kits ca-
    pable of detecting arsenic water concentrations close
    to 10 ug/L. Environmental Science and Technology 40:
    3362-3366 (2006).
Su, C. and R.W. Puls. Arsenate and arsenite removal by
    zerovalent iron: Kinetics, redox  transformation,  and
    implications for in situ groundwater remediation. En-
    vironmental Science and Technology 35: 1487-1492
    (2001).
Swartz,  C.H., N.K.  Blute,  B.  Badruzzman, A. AN, D.
    Brabander, J. Jay, J. Besancon, S. Islam, H.F. Hemond,
    and C.F. Harvey.  Mobility of arsenic in a Bangladesh
    aquifer: inferences from geochemical profiles, leaching
    data, and mineralogical characterization. Geochimica
    et Cosmochimica Acta 68: 4539-4557 (2004).
Tadanier, C.J., M.E. Schrieber, and J.W. Roller. Arsenic mo-
    bilization through microbially mediated deflocculation
    of ferrihydrite. Environmental Science and Technology
    39:3061-3068(2005).
Tallman, D.E. and A.U. Shaikh. Redox stability of inorganic
    arsenic(lll) and arsenic(V) in aqueous solution. Analyti-
    cal Chemistry 52: 199-201 (1980).
USEPA. Ambient Water Quality Criteria for Arsenic -1984.
    U.S. Environmental Protection Agency, Regulations and
    Standards, Criteria and Standards Division, Washing-
    ton  DC (http://www.epa.gov/ost/pc/ambientwqc/arse-
    nic1984.pdf) (1985).
USEPA. Contaminants and remedial options at selected
    metal-contaminated sites, EPA/540/R-95/512, Wash-
    ington DC: Office of Research and Development,  U.S.
    Environmental Protection Agency (1995).
USEPA.   Technology Alternatives  for the Remediation
    of Soils Contaminated with As,  Cd, Cr, Hg, and Pb,
    EPA/540/S-97/500, Washington  DC: Office of Emer-
    gency and Remedial  Response, U.S.  Environmental
    Protection Agency (1997).
USEPA. Drinking Water Standard for Arsenic, EPA 815-
    F-00-015. Washington DC: Office of Water, U.S. Envi-
    ronmental Protection Agency (2001).
USEPA. Proven Alternatives for Aboveground Treatment
    of Arsenic in Groundwater, EPA-542-S-02-002. Wash-
    ington DC:  Office of Solid  Waste and Emergency
    Response,  U.S.  Environmental Protection  Agency
    (2002).
USEPA. 2006 Edition of the Drinking Water Standards and
    Health Advisories. EPA 822-R-06-013, Office of Water,
    Washington DC, (2006a).
USEPA. National Recommended Water Quality  Criteria,
    4304T, Office of Water, Washington DC (2006b).
Vroblesky, D.A. and F.H. Chapelle. Temporal and spatial
    changes of terminal electron-accepting processes in
    a petroleum hydrocarbon-contaminant aquifer and the
    significance for contaminant biodegradation.  Water
    Resources Research 30: 1561-1570 (1994).
Webster, J.G. The solubility of As2S3 and speciation of As
    in dilute and sulphide-bearing fluids at 25 and 90 C.
    Geochimica et Cosmochimica  Acta 54:  1009-1017
    (1990).
Wegrzynek, D. and B. Holynska. Simultaneous analysis of
    trace concentrations of  lead and arsenic  by energy-
    dispersive X-ray fluorescence spectrometry. Applied
    Radiation and Isotopes 44: 1101 -1104 (1993).
Welch, A.M., D.B. Westjohn, D.R. Helsel, and R.B. Wanty.
    Arsenic in ground water of the United  States: Occur-
    rence  and geochemistry. Ground Water 38: 589-604
    (2000).
Wilkie, J.A. and J.G. Hering.  Rapid oxidation of geother-
    mal arsenic(lll) in streamwaters of the eastern Sierra
    Nevada.  Environmental Science and Technology 32:
    657-662  (1998).
Wilkin, R.T. Mineralogical Preservation of Solid Samples
    Collected from Anoxic Subsurface  Environments.
    USEPA Ground Water Issue Paper, EPA/600/R-06/112,
    Cincinnati, OH: U.S. Environmental Protection Agency
    (http://www.epa.gov/ada/download/issue/600R06112.
    pdf) (2006).
Wilkin, R.T. and R.G. Ford.  Use of hydrochloric acid for
    determining solid-phase arsenic partitioning in sulfidic
    sediments.  Environmental Science  and Technology,
    36:4921-4927(2002).
Wilkin, R.T. and R.G. Ford.  Arsenic  solid-phase partition-
    ing in reducing sediments of a contaminated wetland.
    Chemical Geology 228: 156-174 (2006).
Wilkin, R.T,  R.D. Ludwig, and R.G. Ford.  Workshop on
    Monitoring Oxidation-Reduction Processes for Ground-
    water Restoration, EPA/600/R-02/002, Cincinnati, OH:
    U.S. Environmental Protection Agency (2002).
Wilkin, R.T, D. Wallschlaeger, and R.G. Ford. Speciation
    of arsenic in sulfidic waters. Geochemical Transactions
    4: 1-7(2003).
Wolthers, M., L. Charlet, C.H. van Der Weijden, PR. van
    der Linde, and D. Rickard.  Arsenic mobility in the am-
    bient sulfidic environment: Sorption of arsenic(V) and
    arsenic(lll) onto disordered mackinawite. Geochimica
    et Cosmochimica Acta 69: 3483-3492 (2005).
Woolson, E.A., J.H. Axley, and PC. Kearney. Comparison
    of a colorimetric  and a coulometric method for  the
    determination  of arsenic in soil  digests. So/7 Science
    111: 158-162 (1971).
Yalcin, S. and X.C. Le. Speciation of arsenic  using solid
    phase extraction cartridges. Journal of Environmental
    Monitoring 3: 81 -85 (2001).
Yang, H.C., J.H. Kim, W.Z. Oh, H.S. Park, and Y.C. Seo.
    Behavior  of hazardous  and radioactive  metals  in a
    laboratory furnace and  a demonstration-scale incin-
                                                    69

-------
    erator. Environmental Engineering and  Science 15:
    299-311  (1998).
Zobrist,  J., RR. Dowdle, J.A. Davis, and  R.S. Oremland.
    Mobilization of arsenite  by dissimilatory  reduction of
    adsorbed arsenate. Environmental Science and Tech-
    nology 34: 4747-4753 (2000).
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                                               Selenium
                         Chunming Su, Robert G. Ford, Richard T. Wilkin
Occurrence and Distribution

Selenium contamination can be derived from weathering of
natural deposits as well as discharge from mining, smelt-
ing, and coal/oil combustion. Most soils contain between
0.1 to 2 mg kg'1, but soils originating from the Upper Cre-
taceous marine sedimentary rocks (shale) show elevated
concentrations of  selenium in about 80,000 km2 of land
in the 17 western  states of the United States (Nolan and
Clark, 1997). In California, shale materials in the Coastal
Ranges along the entire western margin of the San Joaquin
Valley contain excessive and varied amounts of selenium
with median values as high as 6.5 to 8.7 mg kg"1 (Presser
and Barnes, 1984). Complex interactions among irrigated
agricultural practices and physical and chemical processes
have contributed to high  selenium concentrations (20 to
1000 ug L1) in ground water underlying low-altitude agri-
cultural areas of the western San Joaquin Valley (Deverel
etal., 1984;  Seller, 1997).

Incidences of waterfowl mortality and reproductive failure
at Kesterson Reservoir, San Joaquin Valley, California was
discovered in 1983 and was attributed to elevated selenium
concentrations in  irrigation  drainage (Ohlendorf et al.,
1986). More follow-up studies have revealed that elevated
selenium concentrations in  water from irrigation  projects
in the Western USA are largely caused by application of
irrigation water to soils derived from shales of Cretaceous
age(Naftz, 1996).

There are 21 former uranium mill sites designated under
the Uranium Mill Tailings Radiation Control Act (Federal
Register, 1995) that show ground-water contamination with
toxic elements including selenium,  arsenic, manganese,
molybdenum, uranium, vanadium, and zinc. Contamination
of  groundwater with selenium  also occurs from surface
coal mining in Wyoming (Blaylock et al., 1995). Waste rock
from phosphorite  mining  in southeastern Idaho contains
selenium, cadmium, vanadium, and other metals. Selenium
concentrations in the hundreds of parts per million  have
been found in soils, plants, and seeps in close proximity
to some mine waste dumps. Mortality of livestock due to
selenosis poisoning has been linked to the contamination of
water and vegetation by mine waste (Piper et al., 2000).

Plume Characteristics

The general patterns in subsurface selenium transport that
may be anticipated under differing geochemical conditions
are represented by a series of field investigations that have
been conducted in aquifers with redox chemistry grading
from oxic to strongly reducing. For oxic conditions (8 mg L1
O2, pH ~5, no detectable Fe2+or sulfide), the field study
conducted by Kent et al. (1995) demonstrated that selenium
(as selenate) transport will likely be dominated by adsorp-
tion reactions with aquifer sediment. While adsorption will
tend to retard selenium transport, this process will not likely
prevent plume expansion.  For  mildly  reducing conditions
(0.032 mg L1 O2, pH ~6, 17-28  mg L1 Fe2+, <0.034 mg L1
sulfide), Kent et al. (1994) presented results that indicated
little selenate attenuation due to insufficient conditions to
support either biotic or abiotic reduction of selenate to less
mobile forms.  In contrast to selenium transport behavior
within the oxic zone of this aquifer, there was little retarda-
tion of selenium transport relative to a conservative tracer
(bromide). In the absence of conditions insufficient for sele-
nium reduction to Se(0) or more reduced forms, the higher
pH and greater concentrations  of anions that compete for
adsorption sites on aquifer minerals resulted in conserva-
tive transport of selenate within the suboxic portion of the
aquifer. For highly reducing conditions (<0.006 mg L1 O2,
pH ~7, <25 mg L1 Fe2+,  <0.5 mg  L1 sulfide), field studies
conducted in the shallow aquifer underlying a wetland sys-
tem at the Kesterson Reservoir, California demonstrated
significant attenuation of selenium (White et al., 1990;
Benson et al., 1991; White et al.,  1991). This attenuation
apparently  resulted from  reduction  of selenate and/or
selenite to Se(0) or more reduced forms upon interaction
with biotic/abiotic components of the aquifer sediments in
the reducing ground water.  These observations generally
indicate that the attenuation of selenium will be correlated
with the extent of its reduction within an aquifer.

Remedial Technologies

Active removal  of selenium from ground water may be
achieved either through  above  ground or subsurface (in-
situr) treatment processes. Examples of these approaches
include 1) above-ground  treatment operations where sele-
nium is removed via adsorption, ion exchange, or chemical
reduction,  2) phytoremedation  for shallow ground-water
systems either through crop planting  or the construction
of wetland systems that intercept ground-water discharge,
and 3) installation of permeable reactive barriers in the
subsurface to intercept a migrating plume.

For above ground treatment operations, layered double
hydroxide precipitates have been  used as sorbents for re-
moving both selenate and selenite from water (O'Neil etal.,
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1990). Magnesium-aluminum and zinc-aluminum layered
double hydroxides show high affinity for both selenate and
selenite (You et al., 2001) with uptake of these anions via
exchange with structural interlayer anions. Chemical reduc-
tion, primarily through stimulation of microbial processes,
has been employed for selenium removal from contaminated
water. An algal bacterial selenium removal facility was de-
signed and constructed at the Panoche Drainage District in
California during 1995-1996 to remove selenium and nitrate
from drainage water (Green et al., 2003). Mass removals of
total soluble selenium of 77% or greater were achieved over
a three-year period. Selenate was removed by assimiliatory
and dissimiliatory bacterial reduction. Laboratory tests have
shown the benefit of using rice straw to promote selenate
removal from drainage water through  microbial reduction
of selenate to selenite and subsequently to colloidal Se(0)
(Zhang  and  Frankenberger, 2003).  In another study, a
flow-through wetland system was constructed consisting
of ten unlined cells that  were flooded continuously. The
global mass balance showed that on average 59%  of the
total inflow selenium was retained within the wetland cells,
whereas selenium outputs included outflow (35%),  seep-
age (4%), and volatilization (2%) (Gao et al.,  2003a). The
major selenium  removal mechanism was the reduction of
selenate to Se(0) and immobilization into the organic phase
of the sediments (Gao et al., 2003b). However, it has been
noted that in some wetland systems, biological volatiliza-
tion of selenium may account for 10-30% of the selenium
removed from contaminated water (Allen, 1991; Hansen et
al.,  1998; Zhang and Frankenburger, 1999).

The  use of phytoremediation as a treatment technology
involves  removal of  selenium from contaminated  water
either through absorption into the plant mass or volatiliza-
tion following uptake and conversion  to organic species
(Banuelos et al.,  2002).  Potential crops tested for the
phytoextraction  of selenium in  central California include
two moderate selenium accumulators,  Indian mustard and
canola (Banuelos et al., 1993). Selenium hyperaccumula-
tors have been identified,  e.g., Astragalus bisulcatus, which
can accumulate up to 0.65% (w/w) selenium. Pickering et
al. (2003) have identified high concentrations of the nonpro-
tein amino acid Se-methylseleno-cystein in young leaves of
this hyperaccumulator, but in more mature leaves, selenate
predominates. Another selenium hyperaccumulator, Stan-
leya pinnata (Brassicaceae) is a  perennial that responded
favorably to  repeated cutting in the greenhouse, a trait
that could prove valuable in field-scale phytoremediation
(Parker  et al., 2003).

The use of zerovalent iron [Fe(0)] for removal of selenium in
aqueous systems has been demonstrated in both laboratory
and field applications. Earlier studies show that selenate
can be reduced to a lower oxidation state by metallic  iron in
water (Baldwin et al., 1981). Murphy (1988) reported that
selenate may be removed by Fe(0) through the formation
of Se(0). Selenium in the form of selenocyanate (SeCN")
in oil refinery wastewater and artificial wastewater could be
removed using Fe(0) through the formation of Se(0) and
ferrous selenide (FeSe). The possible chemical  reaction
between SeCN" and Fe(0) included deselenation of SeCN"
and electrochemical reduction of Se(0) to Se(-ll) (Meng
et al., 2002).  Morrison et al. (2002) have demonstrated
selenium removal from ground water using Fe(0) in a field
installation  at  a  uranium mill tailings repository.  Three
treatment cells [Cercona foamed Fe(0) plates, steel wool,
and Peerless granular Fe(0)] were operated for passive
removal of selenium along with arsenic,  manganese, mo-
lybdenum, uranium, vanadium, and zinc from ground water
(Morrison et al., 2002).  Significant removal was achieved
in the treatment  cell using Cercona foamed  Fe(0) plates
(reduction from 202 ug L1 to 6 ug L1  dissolved Se), and
the cell performed for more than three years.

Regulatory Aspects

Selenium was  ranked number 144 out of a total of 275 on
the CERCLA Priority List of Hazardous Substances in 2001,
which was prepared by the Agency for Toxic Substances
and Disease Registry, Centers for Disease Control and Pre-
vention. The USEPA has set the maximum contaminant level
(MCL) at 0.05 mg L1 for drinking water (USEPA, 2006a; http:
//www. epa.gov/waterscience/criteria/drinking/dwstandards.
pdf). The USEPA has found selenium to potentially cause
the following health effects when people are exposed to it
at levels above the MCL: hair or fingernail loss, numbness
in fingers or toes, nervous and circulatory systems prob-
lems, and kidney and liver damage. For non-potable water
sources, ambient water quality criteria  (AWQC) that are
protective of aquatic life may  serve as alternative cleanup
goals. For selenium, current statutes list the chronic crite-
rion for fresh waters as 0.005 mg L"1 with recommendation
for an acute criterion pending review (USEPA, 2006b; http:
//www.epa.gov/waterscience/criteria/nrwqc-2006.pdf). An
example of where this  criterion may apply is  a site where
contaminated ground water discharges to surface water.

Selenium (Se) is an essential nutrient  for humans and
vertebrates,  but  is only  required  in small amounts and
has a very  narrow range between deficient  and  toxic
levels (National  Research Council,  1983).   It is  found
in the enzyme glutathione peroxidase  that inhibits the
oxidative role of  peroxides and hydroperoxides, thereby
protecting immunocomponent cells and  slowing  down
aging processes. Selenium also  plays an important role
in anticarcinogenic activity and prevention of heavy metal
toxic effects. Selenium detoxifies  mercury in humans via
possible formation of 1:1 Hg-Se compounds  (Kosta et al.,
1975). Selenium yeast can both prevent the accumulation of
arsenic in the human body and rectify the damages (Wang
et al., 2001). Selenium supports efficient thyroid hormone
synthesis and  metabolism and protects  the thyroid gland
from damage by excessive iodide exposure (Zimmermann
and Kohrle, 2002). Epidemiological studies have indicated
an inverse relationship between selenium intake and the
incidence of certain cancers; blood or  plasma levels of
selenium are usually lower in patients  with  cancer than
those without this disorder (Whanger, 2004).  An endemic
human disease referred to as Keshan disease in certain
regions of China was caused by selenium deficiency (Chen
etal., 1980). Keshan disease is a juvenile cardiomyopathy
that presents as congestive heart failure in infants and
                                                    72

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young children. This disease has been virtually eliminated
by supplying sodium selenite pills to those at risk. Another
selenium deficiency disease has been also  reported in
areas of China and referred to as Kaschin-Beck disease,
which is an osteoarthropathy, a generative articular disease
caused by  oxidative  damage to cartilage  that leads to
deformation of bone structure (Ge and Yang, 1993).

In contrast, excess selenium is very toxic and can cause
selenium poisoning  (selenosis)  in humans and animals.
Consumption of feed containing greater than  5 mg of Se
kg'1 leads to selenium poisoning in animals (Anderson et
al., 1961;  National Research Council,  1980). Selenium
toxicity disorders  in livestock known as "alkali disease"
and "blind staggers" were widely recognized in the 1930s
in South Dakota in areas of high soil selenium (Magg and
Glen, 1967). Sheep grazing in areas of phosphate mining
operations in southeastern Idaho have died as a  result of
high selenium concentrations in forage and water (Fessler
et al., 2003). Selenium toxicity also has been reported in
some regions of Australia as a result of livestock feeding
on selenium accumulative plant species that contain sele-
nocystathionine (Tinggi, 2003). Endemic selenium poison-
ing of humans characterized by loss of nails and hair was
reported in Yutangba of central China in the 1960s, which
was associated with the use of carbonaceous shales (stone
coal) high in selenium in the form  of elemental selenium
(Zhu et al.,  2004).

Geochemistry and Attenuation Processes

Aqueous Spec/at/on

Selenium is a  metalloid exhibiting  physical  and chemical
properties between that of metals and nonmetals. It chemi-
cally resembles sulfur and exists in organic  and inorganic
chemical forms. Inorganic species include selenide [Se(-ll)],
elemental selenium [Se(0)j, selenite [Se(IV)], and selenate
[Se(VI)]. Organic species include methylated compounds,
selenoamino acids, selenoproteins and their  derivatives.
The speciation of selenium is greatly influenced by the pH
and redox conditions  of the environment (Figure 7.1). For
example, Se(-ll) exists in a reducing environment as hydro-
gen selenide (H2Se) and as metal selenides. Reduction of
selenate to selenite and Se(0) has been shown to decrease
its mobility  in saline, mildly alkaline groundwater (White
et al., 1991). When dissolved in water, H2Se  can oxidize
to elemental selenium. Elemental selenium is stable in a
reduced environment, but it can be oxidized to selenite and
to selenate by a variety of microorganisms (Sarathchandra
and Watkinson, 1981).

Various strains of bacteria have been identified to facilitate
selenate  reduction in soil and  sediment systems. Two
microbial processes, namely methylation of  selenium and
reduction of both  selenate and selenite to Se(0), have a
major influence on  the fate and mobility of this element in the
environment (Dungan and Frankenberger, 1999;  Dungan
et al., 2003). Methylation of selenium,  and subsequent
selenium volatilization, leads to dissipation of soil selenium
to the  atmosphere. Environmental factors such  as the
existing microbial  community, pH, temperature, moisture,
and  organic amendments control the rate of  selenium
volatilization from seleniferous soils  (Frankenberger and
Karlson,  1989; Zhang and Frankenburger, 1999). Under
flooding conditions, part of the methylated selenium may be
transported in water, thus decreasing selenium volatilization
to the atmosphere  (Zhang  and Frankenburger, 1999).
The  addition of organic amendments to soils has been
reported  to stimulate  indigenous microbes to methylate
selenium (Abu-Erreish et al., 1968; Frankenberger and
Karlson,  1989);  whereas,  organic substrates added to
ponded sediments have  been found to  accelerate  the
reduction of selenate and selenite to Se (Tokunaga et al.,
1996) and similar effects were reported in laboratory batch
experiments (Zhang et al., 2003).
   .5
HI
  -.5
     25'
                                 10
                                       12
                                             14
                        PH
Figure 7.1  Eh-pH diagram for selenium at 25 C using
            thermodynamic data from Seby et al. (2001).
            ZSe=10-s (790 ug L1).  Solid stability field for
            elemental selenium is shaded pink.

Mechanisms for selenium reduction by microbes are com-
plex as it occurs under both aerobic (Lortie et al.,  1992)
and anaerobic conditions (Oremland etal., 1989; Oremland
et al.,  1990; Tomei et al., 1992). Both dissimilatory  and
detoxification mechanisms are possible (Oremland, 1994).
The occurrence of sequential reduction of selenate to sel-
enite and then to Se(0) is suggested after amendment of
contaminated soils with barley straw under field capacity
moisture conditions (Camps Arbestain, 1998). The rates
of selenate to selenite reduction  in waste waters is pro-
portional to their respective concentrations in solution and
also to the amount of the microbial  biomass (Rege et al.,
1999). A recent study has documented the occurrence of
both intracellular and  extracellular Se(0) granules in three
phylogenetically and physiologically distinct bacteria that
are able to respire selenium oxyanions, suggesting that
this phenomenon appears to be widespread among such
bacteria (Oremland et al., 2004).The metal sites of selenate
reductase from Thauera selenatis have been characterized
                                                    73

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(Maher et al., 2004); the enzyme was found to contain Se in
a reduced form (probably organic) and the Se is coordinated
to both a metal (probably Fe) and carbon.  Assessment of
Se(IV) and Se(0) reduction in anaerobic  estuarine sedi-
ment microcosms demonstrated the sequential formation of
Se(0) and ultimately Se(-ll) during incubation with elemental
selenium and/or lactate as electron donors for microbial
reduction (Herbel et al., 2003). Dissolved Se(-ll) did not
accumulate in  sediment pore water during incubation due
precipitation with ferrous iron to form FeSe(s), which was
determined via solid phase characterization using X-ray
absorption spectroscopy.

Solubility

Selenium has  the potential of  forming precipitates for all
of its oxidation states (Seby et al., 2001; see Figure  7.2).
For selenate and selenite, this includes precipitates with
common major cations in ground water (Ca, Mg) as well
as transition metals (Fe, Mn) or heavy metals that may be
anticipated in contaminant plumes (Rai et al., 1995; Shar-
masarkar et al.,  1996). These phases are anticipated to
primarily be significant in situations where selenium concen-
trations are highly elevated.  Reduction to form elemental
selenium [Se(0)] can result in very low concentrations of
dissolved selenium. In  general, it has been observed that
selenium reduction to insoluble Se(0) results in immobili-
zation and stabilization of this  element  in the  soil matrix,
since the re-oxidation reaction of Se(0) to soluble selenate
and selenite is relatively slow (Tokunaga et al., 1994; Za-
wislanski and Zavarin, 1996). Abiotic reduction of selenite
to Se(0) was also suggested in sulfate-reducing bacteria
(SRB) biofilms (Hockin  and Gadd, 2003). Elemental sele-
nium and elemental sulfur were found to precipitate outside
SRB cells. Further reduction to selenide [Se(-ll)] can lead
to precipitation of metal selenides, including ferrous iron
and manganese selenides, similar to the formation of metal
sulfides under  sulfate-reducing conditions.

While not commonly observed, it is anticipated that many
suboxic geological environments contain green rust, which
is a mixed  ferrous, ferric hydroxide that  also contains
interlayer anions such as sulfate  and carbonate in its
structure (Feder  et al., 2005).  Identification of green rusts
in sediments and soils is hampered by the rapid oxidation
of green rusts by atmosphere oxygen, and for this reason,
they have not been commonly  reported. However, recent
thermodynamic and spectroscopic studies give direct evi-
dence for the existence of green rusts in soils  (Hansen et
al.,  1994; Trolard et al.,  1997; Feder et al., 2005). Due to
high reactivity,  green rust minerals  are envisioned as po-
tential reducing agents  of a number of contaminants such
as nitrate, chromate, and selenate. Direct evidence for the
formation of reduced selenium species in anoxic sediments
via  abiotic  redox reactions with sulfate green rust  was
provided using X-ray absorption near-edge spectroscopy
(XANES) and Fourier-transform extended X-ray absorption
fine structure (EXAFS) spectroscopy (Myneni et al., 1997).
The mechanism of selenate reduction was described by
the following equation:
      HSe04- + 4Fe4"Fe2lll(OH)12S04.:
HSe- + 8Fe3O4
                     4SO42' + 8H+
32H2O
                                                 (2)
in which sulfate green rust was oxidized to form magnetite,
whereas selenate was reduced to Se(0) and subsequently
to selenide (herein referred to as the 'adsorption-reduction'
pathway).  In addition, a laboratory study has demonstrated
that a significant fraction of dissolved selenate can be co-
precipitated with Fe(ll) and Fe(lll) ions to form Fe(ll)-Fe(lll)
hydroxyselenate green rust with simultaneous reduction of
an equal amount of selenate anions to selenite anions (Re-
fait et al., 2000). In the subsurface environment, selenate
reduction by coprecipitation and adsorption pathways can
occur when selenium-contaminated sediment becomes
reducing (Pickering  et al., 1995).  In the coprecipitation-
reduction pathway, reductive dissolution of Fe(lll) oxides
precipitates green rust with selenate followed by selenate
reduction to Se(0) and selenide.  Reduction of selenium
oxyanions to Se(0) has also been observed in the pres-
ence of iron sulfides (Bruggeman et al., 2005) and ferrous
hydroxide  (Zingaro et al.,  1997).
     LU
       -.5
                                   10
                                         12
                                              14
                           PH
Figure 7.2 Eh-pH diagram for selenium at 25 C using
            thermodynamic data from Seby et al. (2001).
            System Se-H2O-Fe-Ca, with LSe=1&5 M
            (790 \jg L1), ZFe=10-4 M (5.6 mg L1), and
            ZCa=10-2 M (400 mg L1). Solid stabil-
            ity fields for elemental selenium, hydrous
            calcium selenite, and ferrous selenide are
            shaded pink.  FeSe2 was suppressed (data
            not available in the Seby et al. review);
            however, the stability field of the diselenide
            would be intermediate between elemental
            selenium and FeSe.

Adsorption

A comprehensive summary of selenate and selenite adsorp-
tion behavior on individual  soil minerals (Fe, Al, and Mn
oxides, kaolinite, and calcite) and whole soils throughout the
United States is documented in Zachara et al. (1994). This
                                                     74

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research demonstrated the importance of Fe and Al oxide
surfaces for selenium adsorption onto aquifer sediments
and highlighted the dependence of the extent of adsorp-
tion on the pH of ground water and the presence of anions
that compete for adsorption sites. Additional review of the
published literature is provided below.

Selenate has been shown to behave like sulfate with mini-
mal adsorption and high mobility (Goldberg and Glaubig,
1988; Neal and Sposito, 1989); whereas, selenite behaves
analogously to  phosphate, with greater adsorption than
selenate (Neal et al.,  1987; Barrow and Whelan, 1989a;
Zhang and Sparks, 1990). Adsorption of selenite on goethite
decreases with increasing pH, with decreasing selenite con-
centration, and with competing anions such as phosphate,
silicate, citrate, molybdate, carbonate, oxalate, and fluoride
(Balistrieri and Chao, 1987). More selenite is adsorbed onto
montmorillonite than on kaolinite (Frost and Griffin, 1977).
Selenite adsorption in seleniferous soils are decreased in
the presence of sulfate, nitrate, and phosphate (Pareek
et al,  2000). Selenite sorption by aluminum hydroxides is
adversely affected by organic acids (Dynes and Huang,
1997). Selenite selectively adsorbs at the carbonate (CO32')
site on calcite (CaCO3) via ionic exchange, forming a two-
dimensional solid-solution of the form Ca(SeO3)x(CO3)1'x at
the interface; under identical chemical conditions, selenate
adsorption  is inhibited (Cheng et  al., 1997). An earlier
study showed selenate substitution in calcite also occurred
(Redder et al.,  1994). Desorption of selenate is faster and
more  nearly complete than selenite after adsorption and
incubation in soil (Barrow and Whelan, 1989b).

Mechanisms of selenium adsorption have been studied from
both macroscale batch and microscale spectroscopic ap-
proaches. The presence of either selenate or selenite lowers
the electrophoretic mobility and decreases the point of zero
charge of amorphous iron hydroxide  (am-Fe(OH)3)and goe-
thite (oc-FeOOH), suggesting inner-sphere complexation for
both selenate and selenite species (Su and Suarez, 2000).
Both in situ attenuated total reflection - Fourier transform
infrared (ATR-FTIR) and diffuse reflectance  infrared Fourier
transform (DRIFT) spectra show bidentate complexes of
selenate with am-Fe(OH)3 and the DRIFT spectra of selenite
on goethite show bridging bidentate complex of  selenite.
These results are consistent with an earlier in situ extended
x-ray  absorption fine structure (EXAFS) spectroscopic
study (Manceau and Charlet,  1994) that shows selenate
forms an inner-sphere binuclear bridging surface complex
on hydrous ferric oxide and goethite. On the contrary,  an
earlier EXAFS study (Hayes et al.,  1987)  concluded that
selenate forms an outer-sphere surface complex on goe-
thite. A recent combined data set of Raman and ATR-FTIR
spectra indicate that both inner- and outer-sphere surface
complexes of selenate occur on goethite, as predominantly
monodentate inner-sphere surface  complexes at pH < 6,
and as predominantly outer-sphere surface complexes at
pH > 6 (Wijnja and Schulthess, 2000).

Site Characterization

The conditions that favor the  mobility of selenium  in the
environment with  respect to adsorption are alkaline pH,
  high selenium concentrations, oxidizing conditions, and high
  concentrations of additional  anions that strongly adsorb.
  Most of the reductive capacity in aquifers resides in the
  sediments (Barcelona and Holm, 1991). It is thus critical
  to assess the predominant ground-water chemistry across
  the plume as well as the biogeochemical characteristics of
  the aquifer sediments that may lead to selenium attenua-
  tion. Table 7.1 provides a summary of potential attenuation
  processes that may be active within an aquifer along with
  a general approach to site characterization to identify the
  active attenuation process(es).

  Aqueous Measurements

  The mobility of selenium depends mainly on the different
  chemical forms in which it is present. Information on the
  chemical speciation of selenium in ground water provides
  part of the context for understanding the processes that may
  control its attenuation within the aquifer. Measurements that
  can be used to assess the chemical speciation of selenium
  include determination of its total dissolved concentration
  along with  quantification of  individual chemical species
  (inorganic or organic) that may occur within the plume.
  Published USEPA methods that may be employed for the
  determination of the total dissolved concentration of sele-
  nium are documented in Table 7.2. Of these methods, only
  the gaseous hydride technique can differentiate between
  inorganic chemical selenium species, namely selenate
  [Se(VI)] and selenite [Se(IV)],  since only selenite forms
  a gaseous  hydride that is detected using this  technique.
  In this  case, selenate may be inferred by the difference
  between measured selenite and total dissolved selenium.
  However, as previously reviewed, there are other inorganic
  (e.g., HSe") and  organoselenium species that  may occur
  within ground water.  Thus, additional speciation methods
  may be needed in order to properly characterize the dis-
  tribution of dissolved selenium species.

  There are several documented approaches to determine
  the chemical speciation of selenium  in water.  These ap-
  proaches can be classified  into two general categories:
  1) those that use a chemical/mass specific detector fol-
  lowing  chromatographic separation of individual aqueous
  species (hyphenated techniques), and 2) those that provide
  direct and indirect detection of individual species prior to or
  following chemical conversion of individual species within
  the sample (e.g., voltammetric techniques). Recent com-
  prehensive reviews are available for hyphenated techniques
  that employ atomic absorption/fluorescence spectrometry
  (Capelo et al., 2006) or inductively coupled plasma-mass
  spectrometry (B'Hymer and Caruso, 2006). For the latter
  technique, Tirez et al. (2000) provide a useful overview of
  possible isobaric (similar ion mass) and polyatomic interfer-
  ences  (due to argon bonding to sample matrix elements
  within plasma) that can bias analytical results for individual
  chemical species of selenium. There are several published
  procedures that employ voltammetry for the detection of
  selenium in  natural water samples (Locatelli and Torsi, 2000;
  Ochsenkuhn-Petropoulou and Tsopelas,  2002; Bertolina
  et al.,  2006).   As outlined by Ochsenkuhn-Petropoulou
  and Tsopelas (2002), not all aqueous selenium species
75

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Table 7.1   Natural attenuation and mobilization pathways for selenium.
   Attenuation Processes
      Mobilization Processes
       Characterization Approach
 Biotic (microbial) or abiotic
 reduction by reduced Fe-
 bearing minerals of sel-
 enate/selenite to elemental
 Se(0) and/or selenide with
 precipitation as a metal
 selenide.
Cessation of microbial processes via
changes in nutrient supply within the
ground-water flow path; reduction
in the mass of reduced Fe-bearing
minerals within the aquifer due to
changes in redox conditions or supply
of soluble iron. Oxidation of Se(-ll)
or Se(0) upon exposure to oxygen or
other oxidants.
Determination of water chemistry, microbial
populations, and/or sediment mineralogy
linked to selenium reduction along with
spatial and temporal variability of  required
chemical/microbial components relative to
selenium transport pathway(s).  Determi-
nation of total concentration and selenium
chemical speciation in aquifer sediments.
 Precipitation of metal sel-
 enates or selenites
Dissolution of metal selenates/sel-
enites due to increased pH
Evaluation of selenium speciation in ground
water.  Determination of total Se in the solid
matrix and suspected components in arse-
nic-bearing mineral.  Evaluation of mineral
solubility relative to ground-water chemistry
and published solubility constants.
 Adsorption to iron oxyhydrox-
 ides, iron sulfides, or other
 mineral surfaces
Desorption at high pH for oxyhydroxides
and sulfides; adsorption inhibition in the
presence of DOC or competing anions.
Reductive dissolution of iron hydroxides
or oxidative dissolution of iron sulfides.
Evaluation of selenium speciation in the aque-
ous phase.  Determination of total Se in the
solid  matrix.  Evaluation  of  selenium solid
phase partitioning using single or sequential
extraction methodologies.  Batch and column
testing to determine Se uptake capacity of
site-specific aquifer  materials with variable
geochemical conditions.
Table 7.2   Published USEPA methods for determination of selenium in aqueous samples.
Method Name and Number
Inductively Coupled Plasma-Atomic Emission
Spectrometry, 601 OB, Revision 2
Inductively Coupled Plasma-Mass Spectrometry,
6020, Revision 0
Selenium (Atomic Absorption, Furnace Tech-
nique), 7740, Revision 0
Selenium (Atomic Absorption, Gaseous Hydride),
7741 a, Revision 1
Selenium (Atomic Absorption, Borohydride Reduc-
tion), 7742, Revision 0
Source
December 1996
http://www.epa.gov/epaoswer/hazwaste/test/pdfs/6010b.pdf
September 1994
http://www.epa.gov/epaoswer/hazwaste/test/pdfs/6020.pdf
September 1986a
September 1994
http://www.epa.gov/epaoswer/hazwaste/test/pdfs/7741a.pdf
September 1994
http://www.epa.gov/epaoswer/hazwaste/test/pdfs/7742.pdf
  Available electronically at http://web1.er.usgs.gov/nemi/method_summary.jsp?param_method_id=5239.
                                                      76

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are electrochemically active [e.g., selenate and (CH3)2Se].
Thus, this analytical approach requires the development and
testing of sample preparation schemes that can separate
individual  electrochemically  active and inactive species
and convert the latter to active forms. Given this analytical
limitation,  a potential advantage of this approach is that
the instrumentation can be deployed in the field setting,
facilitating rapid screening analyses.

It is important to insure that the in-situ speciation of aque-
ous selenium  is  maintained following sample collection
prior to analysis.  Gomez Ariza et al. (2000) provide a
recent  review of studies  conducted to assess appropriate
methods for selenium species preservation in natural water
samples. In general, this review suggests that acidification
followed by sample storage at <4 C is adequate to prevent
changes in the oxidation state of selenite/selenate and/or
selenium loss  via coprecipitation with ferric (hydr)oxides
due to air exposure over reasonable time periods. More
stringent collection, handling, and storage  requirements
may be necessary for ground-water samples with significant
concentrations of organoselenium species.  Alternatively,
it may  be  possible to carry out separation of the various
selenium species within the field using solid phase extrac-
tion (SPE) immediately following sample collection. There
are several published reviews of the use of solid phase
extraction media that may be employed to selectively extract
individual selenium species from water samples (Bueno et
al., 2002; Camel, 2003; Wake et al., 2004).  Subsequent
analysis of the  retained  selenium species can then be
achieved in the laboratory following elution of the species
off the SPE material.  This approach may be  necessary
for water samples containing selenium species that are
not amenable  to chemical preservation within the native
ground-water sample.

Solid Phase Measurements

Determination of the chemical speciation of selenium within
aquifer sediments may be required for determination of the
mechanisms active in its sequestration from ground water.
This analysis may include determination of the oxidation
state of selenium within the whole solid matrix or individual
components, as well as the determination of its presence
as a discrete  solid phase (e.g., elemental selenium) or
its association with specific mineral components (e.g.,
iron oxides or  sulfides).  The following review provides a
summary of approaches that have been used to assist in
identification of  the solid phase associations  controlling
selenium solubility within a contaminant plume.

There are several examples within the literature of the de-
termination of the distribution of selenium oxidation states
within a solid matrix following partial or complete dissolution
of the solid (e.g., Ferrietal., 1998; Ochsenkuhn-Petropou-
lou and Tsopelas, 2002; de Leon et al., 2003).  However,
there has been insufficient assessment of the potential for
selenium species transformation that may occur with use
of aggressive extraction procedures designed to achieve
complete sample dissolution (especially oxidation state)
in order to recommend this  analytical approach.  Deter-
mination of selenium oxidation state for extracts that are
non-selective for various mineralogical associations within
the aquifer sediment may not provide sufficient detail to
identify  specific  partitioning mechanisms.  In these  in-
stances, the use of extraction solutions that are intended
to target specific mineralogical associations may provide a
necessary compliment to determination of the distribution
of selenium oxidation states within the whole solid matrix.
Again, there are numerous examples of the use of single
or sequential extraction procedures to determine the min-
eralogical speciation of selenium in soils and sediments.
The following provides a synopsis of the most commonly
employed procedures.

Sequential  extraction procedures  (SEP)  employed for
characterization of aquifer sediments can be categorized
into two types: 1) procedures designed to quantify the
abundance of various mineralogical components (e.g., iron
oxides,  carbonates, organic matter), and 2)  procedures
designed to  quantify the abundance of  specific trace ele-
ment phase associations. The first SEP types have a longer
history of development and application and are considered
as an appropriate supplement to other analytical methods
(e.g., X-ray diffraction, thermal analysis) used to define soil/
sediment mineralogy. There are examples in the literature
where this type of SEP has been used to infer selenium
speciation in soil/sediments (e.g., Tokonaga et al., 1991).
However, the chemical extractants employed in these pro-
cedures are not designed to specifically recover selenium
from specific solid phase associations.  In the  absence of
suitable  standard  reference  materials  (soils/sediments)
and the general failure to include matrix spike analyses
with reference compounds that represent various selenium
solid phase  species anticipated in soils/sediments (e.g.,
selenate/selenite adsorbed to reference soil minerals,  el-
emental selenium, organoselenium compounds), the use of
this approach is not likely to provide useful characterization
information.

Sequential extraction procedures designed to target specific
selenium solid phase associations (i.e., type two  SEPs) have
been developed and applied for characterization of soils and
sediments.  A recent study by Wright et al. (2003) provides
a comprehensive review and analysis of the utility of these
procedures.   Ultimately, these authors  demonstrated the
non-selectivity of existing SEPs and provided analysis of an
alternative SEP designed to target selenium in the following
solid phase associations: 1) weakly adsorbed selenate, 2)
strongly adsorbed selenite,  3) elemental selenium [Se(0)],
and 4) selenium associated with the organic fraction. The
authors  provide a comprehensive analytical assessment
of potential analytical artifacts and  extraction phase se-
lectivity  through the use of model reference compounds
incorporated  into either  a 'synthesized' reference soil or
spiked into real soils/sediments for analytical performance
analysis. Wright et al. (2003) point out that all  procedures
are not  capable of uniquely identifying selenium in either
an iron-selenide precipitate (e.g., FeSe,s)) or selenium  in-
corporated into iron sulfides (e.g,, mackmawite or pyrite).
Velinsky and Cutter (1990) suggest the  use of extractants
                                                     77

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designed to target dissolution of iron sulfides as a means
for assessing the amount of coprecipitated selenium, but
this avenue requires additional testing to properly assess its
reliability. Chu et al. (2006) also point to the potential use
of carbon disulfide as a stand-alone extractant to target the
most labile form of elemental selenium.  Finally, it is also
important to circumventing  possible analytical artifacts at
the stage of selenium detection in the extraction solution
due to interferences from either 1) reagents employed in the
extraction and sample pre-treatment for analysis (Wright et
al., 2003) or 2) sample matrix components that were liber-
ated from the soil or sediment (Bujdos et al., 2000).
Recently, X-ray absorption  spectroscopy/microscopy em-
ploying synchrotron radiation sources have been employed
to provide direct assessment of the in-situ speciation of
selenium in soils/sediments. The synchrotron X-ray fluo-
rescence microprobe (SXRFM) provides the capability of
mapping elemental distributions at a spatial  resolution
which, in some cases, approaches 1 urn. The application of
SXRFM for mapping selenium distribution within both natu-
ral and  laboratory-constructed soil aggregates illustrates
the potential usefulness of this tool for site characterization
(Tokunaga et al., 1994;  Strawn et al.,  2002). In addition,
X-ray absorption spectroscopy has  been employed to iden-
tify the in-situ chemical speciation of selenium (including
selenium oxidation state and bonding environment) in both
experimental and natural materials (Tokunaga et al.,  1994;
Strawn et al., 2002; Herbel et al., 2003; Templeton et al.,
2003).  The advantage of  these techniques is the ability
to characterize the solid phase speciation of selenium at
relatively low concentrations in soils/sediments with mini-
mal sample handling that may perturb the characteristics
of the as-collected  sample. However, these methods are
not routinely available, which limits their utility as a regular
component of a site characterization effort.

Selenium Isotope Fractionation
A review of the use of selenium isotope measurements for
the determination of selenium reduction processes active
within an aquifer or the determination of selenium sources
within a plume  is provided in Johnson  and Bullen (2004).
The six stable isotopes, 74Se, 76Se, 77Se, 78Se, 80Se, and
82Se, are present in abundances of 0.889, 9.366, 7.635,
23.772, 49.607, and 8.731%. Reduction of selenate and
selenite, either through  biotic or abiotic processes,  may
result in enrichment of  lighter isotopes in the reduction
products and a complementary enrichment of the heavier
isotopes in the remaining unreduced selenium due to kinetic
isotope fractionation processes (Johnson and Bullen, 2004).
Since reduction of selenium results in breaking a bond with
oxygen, the kinetic effect is a result of the smaller expense
of energy by the microbe to reduce the lighter isotope. The
current state of knowledge relative to the extent of selenium
isotope fractionation as a result of abiotic-biotic oxidation/
reduction processes is summarized in Table 7.3 (Johnson
and Bullen, 2004). This summary is in part derived from
recent studies designed to assess the importance  of biotic
(Herbel et al., 2000; Ellis et al., 2003) and abiotic (Johnson
and Bullen, 2003) reduction processes that are anticipated
to be active  in reducing subsurface environmentals. Cur-
rently available data indicate that reduction of the selenium
oxyanions, selenate and selenite, is the main  source of
fractionation observed in ground-water systems.

Determination of selenium isotope fractionation can be car-
ried out using either thermal ionization mass spectrometry
(TIMS; Johnson et al., 1999) or multi-collector inductively
coupled plasma mass spectrometry (MC-ICP-MS; Rouxel
et al., 2002). Both methods are susceptible to measure-
ment bias due to changes in selenium speciation and mass
distribution that may occur during sample preparation due
to the susceptibility of selenium to redox transformations
and volatilization.  Therefore, it is  recommended that a
'double spike' technique be employed to correct for these
analytical inaccuracies (Johnson  et al.,  1999; Johnson
and Bullen, 2004). With this method, a reference sample
with known 82Se/74Se ratio is added to the sample and its
recovery is monitored along with evaluation of the natural
ratio of ^Se/^Se originally in the sample.  In addition, the
MC-ICP-MS method is also susceptible to bias due to the
formation of  polyatomic species with identical masses that
form within the argon plasma (e.g., ArCI+ and ArC+; Rouxel
et al.,  2002). It should be noted that a representative inter-
laboratory standard is not commercially available, although
the National Institute of  Standards and Technology does
market  a reference meteorite standard (Canyon Diablo
Table 7.3  Review of selenium isotope fractionation ranges for abiotic-biotic processes during reduction and oxida-
           tion (Johnson, 2004).  NM = not measured
Reduction
Se(VI) - Se(IV)
Se(IV) - Se(0)
Se(0) - Se(-ll)
Abiotic
7-12%o
6-13%o
0%0
Biotic
3-5 %o
6-9 %o
NM
Oxidation
Se(IV) - Se(VI)
Se(0) - Se(IV)
Se(-ll) - Se(0)
Abiotic
0%0
0%0
Small
Biotic
NM
NM
NM
                                                     78

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Troilite; SRM 3149).  Typically, a provisional standard is
generated via mixtures of synthetic solutions with known
selenium isotope content, and this reference is  used to
monitor analytical performance.  Isotopic measurements
may be performed on the native ground water or chemical
extracts used to determine the selenium isotope distribution
in on whole or selective fractions of aquifer solids.

Long-term Stability and Capacity

The long-term stability of selenium immobilized onto aquifer
solids will depend on the  prevailing ground-water  chemis-
try over time relative to the conditions that existed at the
time of immobilization.  If there are  significant changes
in ground-water  chemistry following immobilization,  then
the  potential exists for remobilization of selenium  with the
establishment of a new mobile plume. A general example
of this type of situation would be if selenium was originally
immobilized  through partitioning to  mineral sulfides within
aquifer sediments under sulfate-reducing conditions. Sele-
nium associated  with sulfide minerals may be re-mobilized
due to sulfide oxidation if the aquifer were to return to more
oxidizing conditions.  Alternatively,  there is evidence that
elemental selenium may be re-oxidized to selenite/selenate,
resulting  in  re-mobilization  of selenium in ground water
(Tokunaga et al.,  1996; Zawislanski and Zavarin, 1996; Losi
and Frankenberger, 1998). However, as demonstrated by
Dowdle and Oremland (1998) the  oxidation of elemental
selenium is  a slow,  microbially-controlled  process  with
rates that are 3-4 orders of magnitude lower than microbial
reduction of  selenate/selenite.

The capacity to sustain continued removal of selenium from
ground water will be  dictated by the evolution of  ground-
water chemistry  through  time and  the continued  activity
of relevant microbial  processes.  For situations in which
adsorption or abiotic precipitation reactions control sele-
nium attenuation, significant changes in pH or increases
in the concentration of competing anions (e.g., sulfate) in
ground water can cause reversal of adsorption reactions or
dissolution of precipitates (e.g., metal selenates/selenites).
Likewise, a change to more oxidizing conditions can cause
destabilization of reduced mineral  phases (e.g., sulfides)
that may sequester selenium during the original attenua-
tion reaction.  For situations in which microbial reduction
controls selenium attenuation, changes in the availability of
required substrates/nutrients as well as potentially compet-
ing electron acceptors can reduce the capacity of the aquifer
to sustain continued attenuation prior to consumption of the
total mass of selenium within the plume. There is laboratory
and field evidence that increases in ground-water nitrate
can inhibit or slow down selenium reduction  (Oremland et
al.,  1989; Benson, 1998). While these examples are not
exhaustive of all possible situations, they do point to the
importance of establishing a monitoring program that is
designed to  consider the potential impact of changes in
site geochemistry relative to the conditions under which
selenium is attenuated within the aquifer.  Ultimately, the
evolution of  biogeochemical  conditions in the  subsurface
will  dictate the success of the MNA remedy.
Tiered Analysis

Determination of the viability of selenium remediation in
ground water via monitored natural attenuation will depend
upon  proper assessment of contaminant loading to the
aquifer and prevailing geochemistry and mineralogy within
the contaminant plume and the down gradient zone  prior
to the point(s) of compliance. The goal of site assessment
is to  demonstrate the process(es) controlling selenium
sequestration onto aquifer solids and the long-term stabil-
ity of  solid phase selenium as a function of existing and
anticipated ground-water chemistry. The following tiered
analysis structure for site characterization provides an ap-
proach to evaluate candidate sites and define the potential
limitations of MNA as part of a remedy for  ground-water
cleanup.

 Tier I Site characterization under Tier I will involve dem-
onstration that the plume is static or shrinking,  has not
reached compliance boundaries, and does not  impact
existing water supplies.  Once this  is established through
ground-water characterization, evidence is  collected to
demonstrate Se partitioning to aquifer solids within the
plume. Rapid movement of contaminants along preferred
flow paths in the unsaturated and saturated zones can be
induced by  hydrologic events such as  heavy rains (e.g.,
McCarthy et al., 1998; Camobreco et al., 1996).  It will be
important to determine that such hydrogeologic features do
not result in contaminants bypassing zones where natural
attenuation is occurring.  If natural attenuation processes
are active throughout the plume, then an observed increase
in solid phase concentrations within regions of the plume
with higher  aqueous concentrations is anticipated,  e.g.,
near the source term. This field partitioning data may be
supplemented by geochemical modeling that incorporates
measured water chemistry (e.g., pH,  Eh, and  major ion
chemistry) throughout the plume to assess the potential for
solubility control by metal selenates/selenites or elemental
selenium.  Since identification of the chemical speciation of
selenium in water samples and aquifer sediments is critical
towards determining the attenuation mechanism(s), it is rec-
ommended that precautions be taken to preserve selenium
speciation during collection, preservation, and processing
of collected  samples (See recommendations and applica-
tion of these methods in Wilkin et  al., 2002 and Ford et
al., 2005). Identification of active sequestration to prevent
selenium migration in ground-water provides justification
for proceeding to Tier II characterization efforts.

Tier II Under Tier II, the apparent rate and mechanism(s)
of attenuation are determined.  Estimates of a site attenu-
ation  rate(s) can  be assessed via a well transect along
the ground-water flow path.  In addition, time-series  data
may be collected at one or more monitoring points within
the plume. This information will allow assessment of the
relative timescales for contaminant immobilization and fluid
transport and determination of whether remediation objec-
tives can be met within the required regulatory time frame.
In addition, the mechanism(s) for attenuation need to be
identified under this stage of site characterization.  This
                                                     79

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effort may require determination of the chemical speciation
of aqueous and solid phase Se, which may be approached
according to the following scheme:

    1)   Determination of selenium solution speciation
        via direct analytical measurements in combi-
        nation with  speciation  calculations based on
        characterized ground-water chemistry;
    2)   Calculation  of the saturation state of ground
        water  relative to measured aqueous chem-
        istry complimented by the possible  isolation
        of  discrete  Se  mineral  phases via density
        separations (or other schemes) in regions of
        the aquifer with highest solid phase concentra-
        tions;
    3)   Determination of aquifer  mineralogy to deter-
        mine the relative abundance of components
        with documented capacity for Sesorption (e.g.,
        Amonette, 2002);
    4)   Identification of selenium association(s) with
        the various solid phase components of aquifer
        solids through combination of chemical extrac-
        tions with microscopic/spectroscopic confirma-
        tion of phase associations, and;
    5)   Demonstration  of  concurrence between  the
        site conceptual  model  and mathematical
        model(s)  that  describe selenium   removal
        mechanism(s).
It is recommended that identification of selenium chemical
speciation in aqueous and solid matrices be conducted us-
ing samples collected in a manner that preserves the in-situ
mineralogy and speciation of selenium. The demonstration
of concurrence between conceptual and mathematical mod-
els describing selenium transport will entail development
of site-specific parameterization of the chemical processes
controlling selenium solid phase partitioning.

Tier III   Once the  attenuation mechanism(s) has  been
identified for the site, the subsequent characterization effort
under Tier III will involve determination of the stability of
immobilized Se and the capacity of the aquifer to sustain
continued uptake.  It is recommended that the stability of
immobilized Se be tested based on the anticipated evolution
of ground-water chemistry concurrent with plume shrinkage.
For example, changes in ground-water redox potential due
to oxygen intrusion or the occurrence of aerobic microbial
processes can exert a significant influence on Se attenuated
in its elemental form [Se(0)j. Therefore, it is recommended
that sediment leach tests be conducted to characterize the
magnitude of Se mobilization as a function of ground-water
redox chemistry representative of existing  and anticipated
site conditions. This may involve the use of microcosm
tests that stimulate in-situ microbial populations toward the
development of redox conditions considered deleterious for
continued Se immobilization.  It is recommended that  the
capacity for Se uptake onto aquifer solids be determined
relative  to the specific mechanism(s) identified in Tier II.
For example, if site characterization under Tier II indicated
that co-precipitation of Se with iron sulfide due to microbial
degradation of organic compounds coupled with sulfate
reduction  occurs  within the aquifer,  it  is recommended
that the mass distribution of organic carbon, sulfate and
ferrous  iron to  support this reaction within the aquifer
be determined.  This site-specific capacity can then be
compared to Se mass loading within the plume in order to
assess the longevity of the natural attenuation process. If
site-specific tests  demonstrate the stability of immobilized
Se and sufficient capacity within the aquifer to sustain Se
attenuation, then the  site  characterization  effort can prog-
ress to Tier IV.  For cases where  contaminant stability is
sufficient but aquifer  capacity is insufficient for capture of
the entire plume,  then a  determination of the benefits of
contaminant source reduction may be necessary.

Tier IV  Finally,  under Tier IV a monitoring plan is estab-
lished along with  contingency plans in the event of MNA
failure.  It is  recommended that the monitoring plan be
designed to establish both continued plume stability and to
identify changes in ground-water chemistry that may lead
to re-mobilization  of attenuated Se. The specific chemical
parameters to be monitored will include those identified
under Tier III that  may halt Se partitioning to aquifer sedi-
ments and/or result in solubilization of either discrete Se
precipitates or aquifer minerals that sequester Se from
ground water. For example, solution phase parameters
that could alter either  Se precipitation or adsorption include
increases in the concentration of competing anions, such
as sulfate, in combination with changes in ground-water
pH. In contrast, the concentration  of dissolved iron or sul-
fate may indicate  the dissolution of an important sorptive
phase within the aquifer (e.g., reductive dissolution of iron
oxyhydroxides or oxidative dissolution of sulfides). Changes
in these parameters may occur prior to observed changes
in solution Se and, thus,  serve as monitoring triggers for
potential MNA failure. In this instance, a contingency plan
can be implemented  that  incorporates strategies to arrest
possible plume expansion beyond compliance boundaries.
Possible  strategies to prevent plume expansion include
pump and treat operations, installation of reactive barriers
to enhance uptake capacity perpendicular to the direction
of plume advance, or enhancement of natural attenuation
processes within the aquifer through the injection of soluble
reactive components.

References
Abu-Erreish, G.M., E.I.Whitehead, and O.E.Olson.  Evolu-
    tion of volatile selenium from soils. So/7 Science 106:
    415-420(1968).
Allen, K.N.  Seasonal variation of selenium in outdoor
    experimental stream-wetland systems. Journal of En-
    vironmental Quality 20: 865-868 (1991).
Amonette, J.E.    Methods for  determination of mineral-
    ogy and environmental availability. In  So/7 Mineralogy
    with Environmental Applications.  J.B.  Dixon and D.G.
    Schulze (Eds.), Soil Science Society of America Book
    Series; Soil Science Society of America, Inc., Madison,
    Wl, pp. 153-197(2002).
                                                     80

-------
Anderson, M.S., H.W. Lakin, K.C. Beeson, E.F. Smith, and
    E. Thacker.  Selenium  in Agriculture. In Agriculture
    Handbook200, USDA, Washington DC (1961).
Baldwin,  R.A., J.C. Stauter, and D.L. Terrell.  Process  for
    the removal of selenium from aqueous system. U.S.
    Patent 4,405,464 (1981).
Balistrieri, L.S. and T.T. Chao. Selenium  adsorption  by
    goethite. So/7 Science Society of America Journal 51:
    1145-1151 (1987).
Banuelos, G.S., G.E. Cardon, C.J. Phene,  L. Wu, S. Ako-
    houe, and S. Zambruski. Soil boron  and  selenium
    removal by three plant  species. Plant and Soil 148:
    253-263(1993).
Banuelos, G.S., Z.Q. Lin, L. Wu, and N. Terry.  Phytore-
    mediation of selenium-contaminated soils and waters:
    Fundamentals and future prospects. Reviews  on En-
    vironmental Health 17: 291-306 (2002).
Barcelona,  M.J. and T.R. Holm.  Oxidation-reduction ca-
    pacity of  aquifer solids.  Environmental Science and
    Technology 25: 1565-1572 (1991).
Barrow, N.J.and B.R.Whelan.Testing a mechanistic model.
    VII. The effects of pH and of electrolyte on the reaction
    of selenite and selenate with a soil. Journal of Soil Sci-
    ence 40: 17-28(1989a).
Barrow, N.J.and B.R.Whelan. Testing a mechanistic model.
    VIII. The effects of time and temperature of incubation
    on the sorption and subsequent desorption of selenite
    and selenate by a soil. Journal of Soil Science 40: 29-
    37(1989b).
Benson, S.M.  Influence  of nitrate on the mobility and  re-
    duction kinetics of selenium in groundwater systems.
    In Environmental Chemistry of Selenium. W.T. Fran-
    kenberger, Jr. and R.A. Engberg (Eds.), Marcel Dekker,
    Inc., New York, NY, pp. 437-457 (1998).
Benson, S.M., A.F White, S. Halfman, S. Flexser, and  M.
    Alavi.  Groundwater contamination at  the  Kesterson
    Reservoir, California, 1. Hydrogeologic setting and con-
    servative solute transport. Water Resources Research
    27: 1071-1084(1991).
Bertolino, FA., A.A.J. Torriero, E. Salinas,  R. Olsina, L.D.
    Martinez, and J. Raba. Speciation analysis of selenium
    in natural water using square-wave voltammetry after
    preconcentration on actived carbon. Analytica Chimica
    Acta 572: 32-38 (2006).
Blaylock,  M.J., T.A. Twafic, and G.F Vance. Modeling sel-
    enite sorption in reclaimed coal mine soil materials.
    So/7 Science 159: 43-48 (1995).
Bruggeman, C., A. Maes, J. Vancluysen, and P. Vandemus-
    sele.  Selenite reduction in Boom clay:  Effect of FeS2,
    clay minerals and dissolved organic matter. Environ-
    mental Pollution 137: 209-221 (2005).
Bueno, M. and M. Potin-Gautier.   Solid-phase extraction
    for the simultaneous preconcentration of organic (se-
    lenocystine) and inorganic  [Se(IV), Se(VI)]  selenium
    in natural waters. Journal of Chromatography A 963:
    185-193(2002).
Bujdos, M., J. Kubova, and V Stresko. Problems of selenium
    fractionation in soils rich in organic matter. Analytica
    Chimica Acta 408: 103-109 (2000).
Camel,  V  Solid  phase  extraction of trace elements.
    Spectrochimica Acta Part B: Atomic Spectroscopy 58:
    1177-1233(2003).
Camobreco, V.J., B.K. Richards, T.S. Steenhuis, J.H. Peverly,
    and M.B. McBride. Movement of heavy metals through
    undisturbed and homogenized soil columns. So/7 Sci-
    ence 161: 740-750 (1996).
Camps Arbestain, M. Effect of straw amendment and plant
    growth on selenium transfer in a laboratory soil-plant
    system. Canadian Journal of Soil Science 78:187-195
    (1998).
Chen, Y.-W., L. Li,  A. D'Ulivo,  and N. Belzile.  Extraction
    and determination of elemental selenium in sediments
    - A comparative study.  Analytica Chimica Acta 577:
    126-133(2006).
Chen, X.,  G. Yang,  J. Chen, X.  Chen, Z. Wen, and K. Ge.
    Studies on the relations of selenium to Keshan  dis-
    ease.  Biological Trace  Elements Research 2: 91-104
    (1980).
Cheng,  L., PF Lyman, N.C. Sturchio, and M.J.  Bedzyk.
    X-ray standing wave investigation of the surface struc-
    ture of selenite anions adsorbed on  calcite. Surface
    Science 382: L690-L695 (1997).
de Leon, C.A.P, K. DeNicola, M.M. Bayon, and J.A. Caruso.
    Sequential extractions  of selenium soils from Stewart
    Lake: total selenium and speciation measurements with
    ICP-MS detection. Journal of Environmental Monitoring
    5: 435-440 (2003).
Deverel, S.G.,  R.J. Gillion, R. Fujii, J.A. Izbicki, and  J.C.
    Fields. Arial distribution  of selenium and other inorganic
    constituents in  shallow ground water  of the San Luis
    Drain  Service area, San Joaquin Valley, California: A
    preliminary study. U.S. Geological Survey Water Re-
    sources Investigation Report 84-4319 (1984).
Dowdle, PR.  and  R.S.  Oremland.  Microbial oxidation
    of elemental selenium in soil slurries  and bacterial
    cultures. Environmental Science and Technology 32:
    3749-3755(1998).
Dungan,  R.S.  and W.T. Frankenberger,  Jr.  Microbial
    transformations of selenium and the bioremediation
    of seleniferous environments. Bioremediation Journal
    3: 171-188(1999).
Dungan, R.S., S.R. Yates, and W.T. Frankenberger, Jr.
    Transformations of  selenate and selenite by Steno-
    trophomonas maltophilia isolated from a seleniferous
    agricultural drainage pond sediment. Environmental
    Microbiology 5: 287-295 (2003).
Dynes, J.J. and P.M. Huang.  Influence of  organic acids
    on selenite sorption by poorly ordered  aluminum hy-
    droxides. So/7 Science Society of America Journal 61:
    772-783(1997).
Ellis, A.S., T.M. Johnson,  T.D. Bullen, and M.J. Herbel.
    Selenium  isotope fractionation by natural microbial
    consortia. Chemical Geology 195: 119-129 (2003).
                                                    81

-------
Federal Register. Title 40 Code of Federal Regulations Part
    192. Health and  Environmental Protection standard.
    Uranium Thorium Mill Tailings 53, 2854-2871  (1995).
Feder,  R,  F. Trolard, G. Klingelhofer, and G. Bourrie.  In
    situ Mossbauer spectroscopy: Evidence for green rust
    (fougerite) in a gleysol and its mineralogical transforma-
    tions with time and depth. Geochimica et Cosmochimica
    Acta 69: 4463-4483 (2005).
Ferri, T, S. Rossi, and P. Sangiorgio. Simultaneous deter-
    mination of the speciation of selenium and tellurium
    in geological matrices by use of an iron(lll)-modified
    chelating resin and  cathodic stripping voltammetry.
    Analytica Chimica Acta 361: 113-123 (1998).
Fessler, A.J., G.  Moller, PA. Talcott, and J.H. Exon. Se-
    lenium toxicity in sheep grazing reclaimed phosphate
    mining sites.  Veterinary and Human  Toxicology 45:
    294-297 (2003).
Ford, R.G., R.T. Wilkin, K.G. Scheckel, C.J. Paul, F. Beck, P.
    Clark,  and T. Lee.  Field Study of the Fate of Arsenic,
    Lead,  and Zinc at the Ground-Water/Surface-Water
    Interface, EPA/600/R-05/161.  Cincinnati, OH: U. S.
    Environmental  Protection  Agency, (http: //www.epa.
    gov/ada/download/reports/600R05161/600R05161.
    pdf) (2005).
Frankenberger, W.T. Jr. and U. Karlson. Environmental fac-
    tors affecting microbial production of dimethylselenide
    in a selenium-contaminated sediment. So/7  Science
    Society of America Journal 53: 1435-1442 (1989).
Frost, R.R. and R.A.  Griffin.  Effect of pH on adsorption
    of arsenic and selenium from landfill leachate by clay
    minerals. So/7 Science Society of America Journal 41:
    53-56  (1977).
Gao, S., K.K. Tanji, Z.Q. Lin,  N. Terry, and D.W. Peters.
    Selenium removal and mass balance in a constructed
    flow-through wetland system. Journal of Environmental
    Quality 32: 1557-1570 (2003a).
Gao, S., K.K.Tanji, D.W. Peters, Z.Q. Lin, and N.Terry. Se-
    lenium removal from irrigation drainage water flowing
    through constructed  wetland cells with  special atten-
    tion to accumulation in sediments. Water, Air,  and Soil
    Pollution 144: 263-284 (2003b).
Ge, K. and G. Yang. The epidemiology of  selenium defi-
    ciency in the etiological study of endemic diseases
    in China. American Journal of  Clinical Nutrition and
    Supplement 57: 259S-263S (1993).
Goldberg,  S. and R.A. Glaubig. Anion sorption on a cal-
    careous, montmorillonitic soil-selenium. So/7  Science
    Society of America Journal 52: 954-958 (1988).
Gomez Ariza, J.L., E. Morales,  D. Sanchez-Rodas, and I.
    Giraldez. Stability  of chemical species in environmental
    matrices. Trends in Analytical Chemistry 19: 200-209
    (2000).
Green, F.B., T.J. Lundquist, N.W.T Quinn, M.A.  Zarate,
    I.X. Zubieta,  and  W.J. Oswald. Selenium and nitrate
    removal from agricultural drainage using the AIWPS
    technology. Water Science and Technology 48: 299-
    305 (2003).
Hansen, H.C.B., O.K. Borggaard, and J. Sorensen. Evalu-
    ation of the free energy of formation of Fe(ll)-Fe(lll)
    hydroxide-sulphate (green rust) and its reduction of
    nitrite. Geochimica et Cosmochimica Acta 58: 2599-
    2608(1994).
Hansen, D., PJ. Duda, A. Zayed, and N. Terry. Selenium
    removal  by constructed wetland: Role  of biological
    volatilization. Environmental Science and Technology
    32:591-597(1998).
Hayes, K.F, A.L. Roe, G.E. Brown, Jr., K.O. Hodgson, J.O.
    Leckie, and G.A. Parks. In situ X-ray absorption study of
    surface complexes: Selenium oxyanions on a-FeOOH.
    Science 238: 783-786 (1987).
Herbel, M.J., T.M. Johnson, R.S. Oremland, and T.D. Bullen.
    Selenium stable isotope fractionation during bacterial
    dissimilatory reduction of selenium oxyanions. Geochi-
    mica et Cosmochimica Acta 64: 3701-3709 (2000).
Herbel, M.J., J.S. Blum, R.S. Oremland, and S.E. Borglin.
    Reduction of  elemental selenium  to selenide:
    Experiments with anoxic sediments and bacteria that
    respire Se-oxyanions. Geomicrobiological Journal 20:
    587-602 (2003).
Johnson, T.M. A review of mass-dependent fractionation
    of selenium isotopes and implications for other heavy
    stable isotopes. Chemical Geology 204: 201-214
    (2004).
Johnson, T.M. and T.D.Bullen. Selenium isotope fraction-
    ation during reduction by Fe(ll)-Fe(lll) hydroxide-sulfate
    (green rust). Geochimica et Cosmochimica Acta, 67:
    413-419 (2003).
Johnson, T.M., T.D. Bullen, and PT Zawislanski. Selenium
    isotope ratios  as indicators of selenium sources and
    cycling of selenium: Results from the northern reach
    of San Francisco  Bay. Environmental Science and
    Technology 34: 2075-2079 (2000).
Johnson, T.M., M.J. Herbel, T.D. Bullen, and PT Zawislan-
    ski. Selenium isotope ratios as indicators of selenium
    sources and oxyanion reduction. Geochimica et Cos-
    mochimica Acta 63: 2775-2783 (1999).
Kent,  D.B., J.A. Davis, L.C.D. Anderson, B.A. Rea, and
    T.D. Waite. Transport of chromium and selenium in the
    suboxic zone of a  shallow aquifer: Influence of redox
    and adsorption reactions. Water Resources Research
    30: 1099-1114(1994).
Kent,  D.B., J.A. Davis, L.C.D. Anderson, and B.A. Rea.
    Transport of chromium and selenium in a pristine sand
    and gravel aquifer: Role of adsorption processes. Water
    Resources Research 31: 1041-1050 (1995).
Kosta, L., A.R. Byrne, and V. Zelenko. Correlation between
    selenium  and  mercury in man following exposure to
    inorganic mercury. Nature 254: 238-239 (1975).
Locatelli, C. and G. Torsi. Determination of Se,  As, Cu,
    Pb, Cd, Zn and Mn by anodic and cathodic stripping
                                                    82

-------
   voltammetry in marine environmental matrices in the
   presence of  reciprocal  interference.   Proposal of a
   new analytical procedure. Microchemical Journal 65:
   293-303 (2000).
Lorti, L, W.D. Gould, S. Rajan, R.G.L. McCready, and K.J.
   Cheng. Reduction of selenate and selenite to elemen-
   tal selenium by a Pseudomonas stutzeri isolate. Ap-
   plied and Environmental Microbiology 58: 4042-4044
   (1992).
Losi, M.E. and W.T. Frankenberger. Microbial oxidation and
   solubilization of precipitated elemental selenium in soil.
   Journal of Environmental Quality 27: 836-843 (1998).
Magg, D.D. and M.W.  Glen. Toxicity of selenium: Farm
   animals. In Selenium in Biomedicine. O. H. Muth, J. E.
   Oldfield, and P. H. Heswig (Eds.), Conn. AVI, Westport,
   pp. 127-140(1967).
Maher, M.J., J.  Santini, I.J. Pickering,  R.C. Prince, J.M.
   Macy, and G.N. George. X-ray absorption spectros-
   copy of selenate reductase. Inorganic Chemistry 43:
   402-404 (2004).
Manceau, A. and L. Charlet.  The mechanism of selenate
   adsorption on goethite and hydrous ferric oxide. Journal
   of Colloid and Interface Science 168: 87-93 (1994).
Martens, D.A. and  D.L. Suarez.   Selenium speciation of
   soil/sediment determined with sequential extractions
   and hydride  generation atomic absorption spectro-
   photometry. Environmental Science and Technology
   31: 133-139 (1997).
McCarthy, J.F, K.R. Czerwinski, W.E. Sanford, P.M. Jardine,
   and J.D. Marsh. Mobilization of transuranic radionu-
   clides from disposal trenches by natural organic matter.
   Journal of Contaminant Hydrology 30: 49-77 (1998).
Meng X., S. Bang, and G.R Korfiatis. Removal of selenocya-
   nate from water using  elemental iron. Water Research
   36: 3867-3873 (2002).
Morrison,  S.J., D.R. Metzler, and B.P  Dwyer.   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
   56:99-116(2002).
Murphy, A.P.  Removal of  selenate from water by  chemi-
   cal reduction. Industrial Engineering and Chemistry
   Research 27: 187-191 (1988).
Myneni, S.C.B.,T.K.Tokunaga, G.E. Brown Jr. Abiotic sele-
   nium redox transformations in the presence of Fe(ll,lll)
   oxides. Science 278: 1106-1109 (1997).
Naftz, D. L. Using geochemical and statistical tools to iden-
   tify irrigated areas that might contain high selenium
   concentrations in surface water. U.S. Geological  Survey
   Fact Sheet FS077-96, (1996).
National Research Council. Mineral Tolerance of Domestic
   Animals: Selenium. National Academy Press, Washing-
   ton DC, pp. 302-420 (1980).
National Research Council. Selenium in Nutrition. Revised
   edition. National Academy  Press, Washington DC,
   p. 24(1983).
Neal, R.H. and G. Sposito. Selenate adsorption on aluvial
    soils. So/7 Science Society of America Journal 53: 70-
    74(1989).
Neal,  R.H., G. Sposito, K.M. Holtzclaw, and  S.J. Traina.
    Selenite adsorption on alluvial soils: I. Soil composition
    and pH effects. Soil Science Society of America Journal
    51: 1161-1165(1987).
Nolan,  B.T and M.L. Clark. Selenium in irrigated agricul-
    tural areas of the  western  United States. Journal of
    Environmental Quality 26: 849-857 (1997).
Ochsenkuhn-Petropoulou, M. and F Tsopelas. Speciation
    analysis of selenium using voltammetric  techniques.
    Analytica Chimica  Acta 467: 167-178 (2002).
Ohlendorf, H.M., R.L. Hothem, C.M. Bunck, T.W. Aldrich, and
    J.F. Moore. Relationships between selenium concentra-
    tions and avian reproduction.  In  Transactions of 51st
    North American Wildlife and Natural Resources Confer-
    ences, Reno, NV, 21-26, March, 1986, 330-342. Wildlife
    Management Institute, Washington DC (1986).
O'Neil, G.A.,  J.W. Novak, and E.S.  Martin.   Method for
    removing arsenic or selenium from an aqueous solu-
    tion containing a substantial background of another
    contaminant. U. S. Patent 4, 935,146 (1990).
Oremland,  R.S.  Biogeochemical transformations of se-
    lenium  in anoxic  environments.   In Selenium in the
    Environment. Frankenberger Jr., W. T. and S. Benson
    (Eds.), Marcel Dekker, New York, NY 389-419 (1994).
Oremland, R.S., J.T. Hollibaugh, A.S.  Maest.TS. Presser,
    L.G. Miller, and W.C. Culbertson. Selenate  reduction
    to elemental selenium by anaerobic bacteria in sedi-
    ments and culture; Biogeochemical significance of a
    novel,  sulfate-independent respiration. Applied and
    Environmental Microbiology 55: 2333-2343 (1989).
Oremland, R.S.,  N.A. Steinberg, A.S. Maest, L.G. Miller,
    and J.T. Hollibaugh.  Measurement of in situ rates of
    selenate removal by dissimilatory bacteria  reduction
    in sediments.  Environmental Science and Technology
    24:  1157-1163(1990).
Oremland, R.S.,  M.J. Herbel, J.S. Blum,  S. Langley, T.J.
    Beveridge, P.M.  Ajayan,  T. Sutto, A.V. Ellis, and S.
    Curran. Structural and spectural features of  selenium
    nanospheres produced by  Se-respiring  bacteria.
    Applied and Environmental Microbiology 70: 52-60
    (2004).
Pareek, N., K.S. Dhillon, and S.K. Dhillon. Effect of sulphate,
    nitrate and phosphate ions on adsorption  of selenium
    in seleniferous soils of Punjab. Journal of Nuclear Ag-
    riculture and Biology 29: 167-174 (2000).
Parker, D.R., L.J. Feist, T.W. Varvel, D.N. Thomason, and
    YQ. Zhang.  Selenium  phytoremediation  potential
    of  Stanleya  pinnata.  Plant  and  Soil 249:  157-165
    (2003).
Pickering, I.J., G.E. Brown Jr., T.K. Tokunaga. Quantitative
    speciation of selenium in soils using X-ray absorption
    spectroscopy. Environmental Science and Technology
    29:2456-2459 (1995).
                                                    83

-------
Pickering, I. J., C. Wright, B. Bubner,  D. Ellis, M. W. Per-
    sans, E.Y. Yu, G.N. George, B.C. Prince, and D.E. Salt.
    Chemical form and distribution of selenium and sulfur in
    the selenium hyperaccumulator Astragalus bisulcatus.
    Plant Physiology 131: 1460-1467 (2003).
Piper, D.Z., J.P Skorupa, T.S. Presser,  M.A.  Hardy, S.J.
    Hamilton, M.  Huebner,  and R.A. Gulbrandsen. The
    phosphoria formation at the Hot Springs mine in south-
    erneastern Idaho: A source of selenium and other trace
    elements to surface  water, ground water,  vegetation,
    and biota, U.S. Geological Survey, USGS Open File
    Report 00-050, 73 pp. (2000).
Presser, T.T. and I. Barnes. Selenium concentration in water
    tributary and in vicinity of the Kesterson National Wildlife
    Refuge, Fresno and Merced Counties, California, Sac-
    ramento, CA; U.S. Geological Survey Water Resources
    Investigation Report, 84-4112 (1984).
Rai, D.,  A.R. Felmy, and D.A. Moore. The  solubility prod-
    uct of crystalline ferric selenite hexahydrate and the
    complexation constant of FeSeO3+. Journal of Solution
    Chemistry 24: 735-752 (1995).
Reeder, R.J., G.M. Lamble, J.F Lee, and W.J.  Staudt.
    Mechanism of SeO42' substitution  in calcite: An XAFS
    study. Geochimica et Cosmochimica Acta 58: 5639-
    5646(1994).
Refait, PH., L. Simon, C.  Louis, and J.M.R.  Genin. Reduc-
    tion of SeO42' anions and anoxic formation of iron(ll)-
    iron(lll) hydroxy-selenate green rust.  Environmental
    Science and Technology 34:  819-825 (2000).
Rege, M.A.,  D.R. Yonge, D.P Mendoza,  J.N. Petersen, Y
    Bereded-Samuel,  D.L.  Johnstone, W.A.  Apel, and
    J.M. Barnes.  Selenium reduction by a  denitrifying
    consortium. Biotechnology  and Bioengineering 62:
    479-484(1999).
Rouxel, O., J. Ludden, J. Carginan, L. Marin, and Y. Fouquet.
    Natural variations  of Se isotopic  composition deter-
    mined by hydride generation multiple collector induc-
    tively coupled plasma mass spectrometry. Geochimica
    et Cosmochimica Acta 66: 3191-3199  (2002).
Seby, F, M. Potin-Gautier, E. Giffaut, G. Borge, and O. F
    X. Donard. A critical review of thermodynamic data
    for selenium species at 25C. Chemical Geology 171:
    173-194(2001).
Seller,  R.L. Prediction of lands  susceptible to irrigation-
    induced  contamination  of water.   In  Environmental
    Chemistry of Selenium. W. T. Frankenburger Jr. and R.
    A. Engberg  (Eds.), Marcel Dekker, New York,  NY pp.
    397-418(1998).
Sharmasarkar, S., K.J. Reddy, and G.F.Vance.  Preliminary
    quantification  of metal selenite solubility  in aqueous
    solutions. Chemical Geology 132:  165-170 (1996).
Sposito, G.,  R.H. Neal, K.M. Holtzclaw, D. Bedford, and
    S.J. Traina.  Sorption and desorption of selenite and
    selenate by representative soils of the west side  of
    the San Joaquin Valley.  In UC Salinity/Drainage Task
    Force 1985-86 Tech. Progress Rep. University of Davis.
    18-89 (1986).
Strawn, D., H. Doner, M. Zavarin, and S. McHugo.  Mi-
    croscale investigation into the geochemistry of arsenic,
    selenium, and  iron in soil developed in pyritic shale
    materials. Geoderma 108: 237- 257 (2002).
Su, C. and D.L.  Suarez. Selenate and selenite  sorption
    on iron oxides: An infrared and electrophoretic study.
    So/7 Science Society of America Journal 64:  101-111
    (2000).
Tan, T.T.Y., D. Beydoun, and R. Amal. Photocatalytic reduc-
    tion of Se(VI) in aqueous solutions in UV/TiO2 system:
    Importance of  optimum ratio of  reactants  on TiO2
    surface. Journal of Molecular Catalysis. A: Chemical
    202-73-85 (2003).
Templeton, A.S., T.P Trainor,  A.M. Spormann, and G.E.
    Brown, Jr. Selenium speciation and  partitioning be-
    havior over time within Burkholderia cepacia biofilms
    formed on a-AI2O3 surfaces. Geochimica et Cosmochi-
    mica Acta 67: 3547-3557 (2003).
Tinggi, U. Essentiality  and toxicity of selenium and  its
    status in Australia:  A review. Toxicology Letters 137:
    103-110(2003).
Tirez, K., W. Brusten, S. Van  Roy, N. De Brucker and L.
    Diels. Characterization of inorganic selenium species
    by ion chromatography with ICP-MS detection in micro-
    bial-treated industrial waste water. Journal of Analytical
    Atomic Spectrometry 15: 1087-1092 (2000).
Tokunaga, T.K., D.S. Lipton, S.M. Benson, A.W. Yee, J.M.
    Oldfather,  E.G. Duckart, PW. Johannis, and K.E.
    Halvorsen.   Soil  selenium fractionation, depth pro-
    files and time trends in a vegetated site at Kesterson
    Reservoir.  Water, Air, and Soil Pollution 57/58: 31-41
    (1991).
Tokunaga, T.K.,  S.R. Sutton, and S. Bajt.  Mapping of
    selenium concentrations in soil aggregates with syn-
    chrotron x-ray fluorescence microprobe. So/7 Science
    158:421-434(1994).
Tokunaga, T.K., I.J. Pickering, and G.E. Brown Jr. Selenium
    transformation in ponded sediments. So/7 Science So-
    ciety of America Journal 60: 781-790 (1996).
Tomei, FA., L.L. Barton, C.L.  Lemanski, and T.G. Zocco.
    Reduction of selenate and selenite to elemental sele-
    nium by Wolinella succinogenes. Canadian Journal of
    Microbiology 38: 1328-1333 (1992).
Trolard, F,  J.M.R. Genin, M. Abdelmoula, G. Bourrie, B.
    Humbert, and A. Herbillon. Identification of a green rust
    mineral in  a reductomorphic soil by Mossbauer and
    Raman spectroscopies. Geochimica et Cosmochimica
    Acta6-\: 1107-1111 (1997).
USEPA. Use of Monitored Natural Attenuation at Superfund,
    RCRA  Corrective Action, and Underground Storage
    Tank Sites, Directive Number 9200.4-17P; EPA, Office
    of Solid Waste and Emergency Response, U.S. Govern-
    ment Printing Office: Washington DC (1999).
USEPA. 2006 Edition of the Drinking Water Standards and
    Health Advisories. EPA 822-R-06-013, Office of Water,
    Washington DC (2006a).
                                                    84

-------
USEPA. National Recommended Water Quality Criteria,
    4304T, Office of Water, Washington DC (2006b).
Velinsky, D.J. and G.A. Cutter. Determination of elemental
    selenium and pyrite-selenium in sediments. Analytica
    Chimica Acta 235: 419-425 (1990).
Wake, B.D., A.R. Bowie, E.C.V. Butler, and RR. Haddad.
    Modern preconcentration methods for the determination
    of selenium species in environmental water samples.
    Trends in Analytical Chemistry 23: 491-500 (2004).
Wang, W., L.Yang, S. Hou, J.Tan, and  H. Li. Prevention of
    endemic arsenism with selenium. Current Science 81:
    1215-1218(2001).
White, A.R, M.L. Peterson, and R.D. Solbau. Measurement
    and interpretation of low levels of dissolved oxygen in
    ground water. Ground Water 28: 584-590 (1991).
White, A.R, S.M. Benson,  A.W.Yee, H.A. Wollenberg, and
    S. Rlexer. Groundwater contamination at the Kesterson
    Reservoir, California, 2. Geochemical parameters influ-
    encing selenium mobility. Water Resources Research
    27: 1085-1098(1991).
Whanger, RD.  Selenium and its relationship to cancer: An
    update. British Journal of Nutrition 91:11 -28 (2004).
Wijnja, H. and C.R Schulthess. Vibrational spectroscopy
    study of selenate and  sulfate adsorption mechanisms
    on Re and Al (hydr)oxide surfaces. Journal of Collloid
    and Interface Science 229: 286-297 (2000).
Wilkin, R.T.,  R.D. Ludwig, and R.G. Rord.  Workshop on
    Monitoring Oxidation-Reduction Processes for Ground-
    water Restoration, EPA/600/R-02/002.   Cincinnati,
    OH: U.S. Office of Research and  Development,  U.S.
    Environmental Protection Agency (2002).
Yamada, H.,  Y Kase, M. Usuki,  S. Kajiyama,  and
    K.  Yonebayashi.  Selective  determination  and
    transformation  of  elemental  selenium in  soils.  So/7
    Science and Plant Nutrition 45: 403-408 (1999).
You, Y, G.R Vance, and H. Zhao. Selenium adsorption on
    Mg-AI and Zn-AI  layered double hydroxides. Applied
    Clay Science 20:  13-25 (2001).
Zachara,  J.M., D. Rai, D.A. Moore, G.D. Turner, and  A.R.
    Relmy.  Chemical Attenuation Reactions of Selenium.
    TR-103535, Electric Power  Research Institute,  Palo
    Alto, California, 1994.
Zawislanski, RT. and M. Zavarin.  Nature and rates of sele-
    nium transformations:  A laboratory study of Kesterson
    Reservoir soils. So/7 Science Society of America Journal
    60:791-800(1996).
Zhang, YQ. and J.N.  Moore. Selenium fractionation and
    speciation in a wetland system. Environmental Science
    and Technology 30: 2613-2619 (1996).
Zhang, YQ.  and W.T. Rrankenburger  Jr.   Effects  of soil
    moisture, depth, and organic amendments on selenium
    volatilization.  Journal of Environmental  Quality 28:
    1321-1326(1999).
Zhang, YQ. and W.T. Rrankenburger Jr. Characterization of
    selenate removal from drainage water using rice straw.
    Journal of Environmental Quality 32: 441 -446 (2003).
Zhang, Y.Q., Z.A.Zahir, and W.T. Rrankenburger Jr. Ractors
    affecting reduction of selenate to elemental selenium
    in agricultural drainage water by Enterobactertaylorae.
    Journal of Agricultural and Food Chemistry 51: 7073-
    7078 (2003).
Zhang, PC. and D.L.  Sparks.  Kinetics  of selenate and
    selenite adsorption/desorption at the  goethite/water
    interface. Environmental Science and Technology 24:
    1848-1856(1990).
Zhu, J., W. Zuo, X. Liang, S. Li, and B. Zheng. Occurrence
    of native selenium  in  Yutangba and its environmen-
    tal implications. Applied Geochemistry 19: 461-467
    (2004).
Zimmermann, M.B. and J. Kohrle. The impact of iron and
    selenium deficiencies on iodine and thyroid metabolism:
    Biochemistry and relevance to public health. Thyroid
    12:867-878(2002).
Zingaro, R.A., D.C. Dufner, A.R Murphy,  and C.D. Moody.
    Reduction of oxoselenium anions by iron(ll)  hydroxide.
    Environment International 23: 299-304 (1997).
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                                                 Nitrate
                                    Chunming Su, Robert G. Ford
Occurrence and Distribution

Nitrate (NO3-) is a chemical form of nitrogen that is essential
for plant growth, but potentially toxic to human and animal
life at moderate concentrations.  Nitrate  is  considered to
be the most ubiquitous chemical contaminant in aquifers
throughout the world with continuously observed  increases
in contamination (Spalding and Exner, 1993). Results from
the USGS National Water Quality Assessment Program
showed that in the United States nitrate is detected in 71 %
of ground water samples, more than 13 times as often as
ammonium  (NH4+),  nitrite (NO2~), organic  nitrogen,  and
orthophosphate,  based on a common detection threshold
of 0.2 mg of N L1  (Nolan and Stoner, 2000). According
to  these authors, shallow ground  water (< 5  m  deep)
beneath agricultural land has the highest median  nitrate
concentration (3.4 mg of N L1), followed by shallow ground
water beneath urban land (1.6 mg of N L1) and deeper
ground water in major aquifers (0.48 mg of  N L1). Nitrate
exceeds the maximum contaminant level (MCL), 10 mg of
N L1, in more than 15% of ground-water samples from 4
of 33 major aquifers commonly used as a source of drink-
ing water  (Nolan and Stoner,  2000).  Recent analyses of
ground-water samples from about 1500 domestic drinking
water and public supply wells show that of more than 140
contaminants measured, nitrate most frequently  exceeded
drinking water standards  or human health criteria (Squil-
lace  et al., 2002).

A National Research Council report has indicated that there
are approximately 300-400 thousand nitrate-contaminated
sites in  the United States (National Research Council,
1994). Sources of nitrate and nitrite in ground water include
atmospheric deposition  from fossil fuel burning, runoff from
fertilizer use; leaching from animal wastes  from confined
animal feedlot operations and dairies, septic tanks,  and
sewage;  solid waste disposal (landfills and waste tips);
and erosion  of natural deposits (Puckett,  1995; Nolan and
Stoner, 2000).  Organic nitrogen and ammonia are potential
nitrate sources because they tend to be converted to nitrate
in natural waters.

While nitrate leaching from disturbed forests is a threat to
ground water, this source is small compared to agricultural
and industrial sources (Keeney, 1986). A survey of eastern
U.S.  watersheds showed that the total nitrogen levels in
streams  draining agricultural watersheds  were  fivefold
greater than forested watersheds, while the percentage
of total nitrogen was much greater from  agricultural than
from forested watersheds  (Omernik, 1976).  Since most
native and extensively managed grasslands are nitrogen
deficient, very little nitrate will be available for leaching; also
many of these grasslands are in the semiarid West where
leaching is limited. On the other hand, intensively managed
forage and grazed grasslands (with major nitrogen inputs
from fertilizer and symbiotic nitrogen fixation) may be the
source of considerable ground water nitrate. Grasslands,
like croplands, have annual above-ground biomass cycles
that leave nitrate in the soil profile susceptible to leaching
at times of the year when plant uptake  is minimal, usually
spring and autumn (Keeney, 1986). Agricultural cropland
provides  a large nonpoint  source relative to other nitrate
sources.   Nitrogen  applications  in  amounts  exceeding
the optimum rates are often used to  provide  maximum
economic yields. Excess nitrate in the root zone is leached
at times when the soil is vulnerable to substantial  rainfall
or excessive irrigation. Irrigation combined with high use of
nitrogen fertilizers is the primary source of nitrate in ground
water (Keeney, 1986).

Most of the areas high in ground-water nitrate are west of
the Missouri River where irrigation is a necessity. Aquifers
in highly agricultural areas in the southeastern USA report-
edly are  not contaminated. Vegetative uptake  and deni-
trification in  this warm, wet, carbon-rich environment are
responsible for the natural remediation of nitrate  in shallow
aquifers (Spalding and Exner, 1993).  In the Middle Atlantic
States and the Delmarva Peninsula,  localized contamina-
tion occurs beneath cropped, well-drained soils that receive
excessive applications of manure and commercial fertilizer.
Extensive tile drainage has for the most part prevented a
nitrate problem in the ground water of the Corn Belt states
in that the nitrate-contaminated recharge is diverted by tile
drains and  subsequently  is discharged to surface water
(Spalding and Exner, 1993).

Point-source nitrate contamination includes ammonium
nitrate as explosives residues in mining operations (e.g.,
Johnson  et al.,  2000).  Nitrate in ground water can also
arise from deposits lain down during geological times. Ex-
amples are nitrate found in Pleistocene age loess of semi-
arid southwestern and western central Nebraska (Boyce
et al., 1976) and high levels of geological nitrogen found
in the alluvium beneath the San Joaquin Valley,  California
(Strathouse  et al., 1980; Holloway et al., 1998). A large
reservoir of bioavailable  nitrogen (up to 104  kilograms
of nitrogen per  hectare, as nitrate) has been previously
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overlooked in studies of global nitrogen distribution (Wal-
voord et al., 2003). The reservoir has been accumulating
in subsoil zones of arid regions throughout the Holocene.
Consideration of the subsoil reservoir raises estimates of
vadose-zone nitrogen  inventories by 14 to 71% for warm
deserts and arid shrublands worldwide and by 3 to 16%
globally. Subsoil nitrate accumulation indicates long-term
leaching from  desert  soils,  impelling  further evaluation
of  nutrient  dynamics in xeric  ecosystems. Evidence that
subsoil accumulations are readily mobilized raises concern
about ground-water contamination as a result of land-use
or climate  change.

Ground water, due to its direct connection to surface wa-
ters, can be a significant source of nitrogen to  lakes,  im-
poundments, and estuaries. Increased  nitrogen input from
ground-water sources has caused overgrowth of aquatic
plants  and subsequently depletion of dissolved  oxygen
as plants decay (eutrophication).  Although overlooked in
many nutrient management plans, ground-water discharges
of  nitrate to streams may contribute the majority of  the
total amount of  stream nitrogen (Williamson et al., 1998).
Ground-water contamination at landfill sites by ammonium
can be significant with mean concentrations of about 900
mg NH/-N L1 having been reported for landfill  leachates
in the United Kingdom (Burton and Watson-Craik, 1998).
Ammonium production is a result of amino acid fermentation
during decomposition of organic matter in domestic waste
disposed under anaerobic conditions. Anaerobic conditions
in the waste may persist for many years particularly if  the
landfill  is capped with a low permeability cover to prevent
water infiltration and gas exchange.  Ammonium attenua-
tion in subsoils  and ground water is predominantly due to
cation exchange and/or nitrification (biological  oxidation)
processes (Buss et al., 2004). Currently, more research is
being conducted to improve the understanding of nitrogen
cycling at  the  watershed  level and to determine where
nitrogen attenuation is occurring and to what extent. This
information is needed to  assist the U.S. Environmental
Protection  Agency and States with the general development
of total maximum  daily load (TMDL) for nitrogen, as well
as to help States adopt more appropriate nutrient criteria
on which TMDLs are based.

Plume Characteristics

Nitrate concentration in ground water is variable and  de-
pends  on  interactions among several factors, including
nitrogen loading, soil type, aquifer permeability, recharge
rate, climate, and aquifer oxidation state and pH.  Factors
that  generally increase  nitrate concentrations in ground
water include well-drained soils,  fractured bedrock,  and
irrigation.   Factors that mitigate nitrate contamination of
ground water include poorly drained soils, greater depth
to  ground  water, artificial  drainage systems, intervening
layers of unfractured bedrock, a low rate of ground water
recharge, and anaerobic conditions in aquifers (Nolan and
Stoner, 2000). Vulnerability of ground water to contamina-
tion by nitrate does not depend on any single factor, but on
the simultaneous influence of factors representing nitrogen
loading sources and aquifer susceptibility (Nolan, 2001).
Tools have been developed to assess aquifer susceptibil-
ity to nitrate contamination based on the factors indicated
above (Ceplecha et al., 2004). Research conducted within
the North Carolina  Coastal Plain illustrates the depen-
dence of nitrate transport on the variability of subsurface
hydrogeologic characteristics. The North Carolina Coastal
Plain  consists of varying permeability surface sediments
underlain by relatively impermeable, highly reduced sedi-
ments (Gilliam et al.,  1974).  Nitrate accumulates in the
shallow  ground water in response to agricultural activities,
but because  of the  reduced zones (up to 15 m thick) it
does not reach deep aquifers (Gilliam et al., 1979). It was
found that  if the water table was  artificially raised  using
flashboard  riser-type water-level control structures on tile
mains or outlet ditches, denitrification in the subsoils  could
be enhanced to  reduce  nitrate concentrations in shallow
ground water (Skaggs and Gilliam, 1981).

Remedial Technologies

A recent review of available technologies for remediation
of  nitrate  contamination  in ground  water  has  been
published by the Interstate Technology Regulatory Council
(ITRC, 2000), which  is available at http://www.itrcweb.
org/Documents/EISBD-1.pdf.   Available technologies
include:  1) ex-situ,  pump-and-treat  approaches that
extract nitrate from  recovered  ground water or degrade
nitrate to  innocuous compounds  through above-ground
biotic or abiotic chemical reduction processes, and  2) in-
situ treatment approaches that degrade nitrate within the
plume via phytoremediation, stimulation of native microbial
communities  for denitrification, or the emplacement of
media that achieve  biotic and abiotic chemical reduction
of  nitrate.  Examples of field demonstrations of  biotic
remediation  technologies  for removal  of  nitrate  from
ground water are listed in Table 8.1.  In addition, research
indicates that abiotic processes may be used to remediate
nitrate via  ex-situ or  in-situ approaches.  Technologies
that extract nitrate from water via concentration onto solid
media (e.g., Tezuka et al., 2004) or into a waste  brine
have  the disadvantage of requiring  further treatment or
disposal of the  process waste stream (e.g., Dorsheimer
et  al., 1997). Examples of  chemical reduction of nitrate
via systems with reduced Fe media (Chew and Zhang,
1998, 1999; Alowitz  and Scherer, 2002; Chen et al., 2004;
Su and Puls,  2004; Chen et al., 2005; Mishra and Farrell,
2005; Yang and Lee, 2005)  indicate that the predominant
reaction leads to the production of ammonia-nitrogen,
although the potential exists to control reaction conditions
to favor complete nitrate reduction to dinitrogen gas. There
is evidence that  zero-valent iron may  also stimulate the
growth of subsurface denitrifying bacteria (Gu et al., 2002),
suggesting the use of this material in permeable reactive
barriers  for mixed contaminant plumes  containing nitrate
and other inorganic/organic constituents more amenable
to chemical reduction.

The success  of these treatment approaches will be influ-
enced by the level of knowledge of the ground-water flow
regime as well as the geochemical characteristics within
the plume. For technologies based on nitrate extraction, the

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Cover cropping in the Northern U.S. Corn
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Monitoring network consisting of well transect
from crop field through hill slope, adjacent flood
plain, and into receiving stream; measured
nitrate removal


Nutrient interception by a riparian forest
adjacent to cropland on Delmarva penin-
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leaf compost); horizontal infiltration beds in
vadose zone and vertical wall or buried contain-
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Permeable reactive barriers to intercept
effluent from septic system infiltration
beds, horizontally migrating septic system
plume, farm field drainage water, or a
shallow sand-and-gravel aquifer

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Network of injection and recovery wells to
control ground-water recirculation in subsurface
reaction zone; operational parameters influenced
denitrification rates and formation clogging


Stimulation of in-situ biological denitrifica-
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source such as ethanol or acetic acid













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Conducted push-pull tests in wells to stimulate
and monitor denitrification; injected solutions
consisted of neutralized site ground water made
anoxic by sparging with 80%N2-20%CO2, and
amended with either ethanol or glucose as an
electron donor


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plete reduction


Sand and gravel aquifer at the Massachu-
setts Military Reservation, Cape Cod near
Falmouth, Massachusetts; nitrate plume
from surface disposal of treated sewage










89

-------
presence of common anions in ground water may reduce
treatment performance for engineered systems based on
ion exchange, reverse osmosis, or electrodyolisis.  For
microbial nitrate reduction (ex-situ or in-situ), geochemical
parameters such as low pH and high total dissolved solids,
as well as the presence of aerobic conditions or compet-
ing electron acceptors such as perchlorate or sulfate may
negatively impact the rates and extent of nitrate reduction
that can be achieved. It is important to consider these site-
specific factors when selecting treatment technologies.

Regulatory Aspects

Nitrate was ranked number 216, nitrite  number 212, and
ammonia number 160 out  of a total of 275 contaminants
on the CERCLA Priority List of Hazardous Substances in
2001 (Centers for Disease Control,  1996). The USEPA has
set the MCL for  nitrate at 10  mg of N L1 (USEPA, 2006;
http://www.epa.gov/waterscience/criteria/drinking/dwstan-
dards.pdf).  The USEPA has determined that infants below
the age of six months who drink water  containing nitrate
or nitrite in  excess of the MCL could become seriously ill
and, if untreated, may  die. Symptoms of acute exposure
include shortness of breath and blueness of the skin, or
blue-baby syndrome (methemoglobulinemia). This is due to
the conversion of nitrate to nitrite by the  body, which inter-
feres with the oxygen-carrying capacity of the child's blood.
Chronic exposure to high levels of nitrate/nitrite results in
diuresis, increased starchy deposits, and hemorrhaging of
the spleen (http://www.epa.gov/safewater/contaminants/
d w_co nta mf s/n it rates. ht m I).

Geochemistry and Attenuation  Processes

Aqueous Spec/at/on

Nitrate (with an oxidation number of +5 for nitrogen) is the
conjugate base of HNO3, a strong mineral acid. Common
metal nitrate salts are highly soluble in water. Nitrate forms
weak coordination complexes with  metals  and displays
minimal sorption to inorganic or organic surfaces. Being an
anion, nitrate (NCy) in solution is attracted electrostatically
to positively charged surfaces and  is repulsed by negative
surfaces. Nitrate is not appreciably adsorbed onto aquifer
sediments.
                                 Redox Chemistry
                                 Denitrification refers to a  microbial respiratory process
                                 where nitrate is used as a terminal electron acceptor and
                                 is reduced to dinitrogen gas by the following generalized
                                 half reaction:
                                  2NCy  +
12
               10e
       6H2O
(Eq.8.1)
                                 It is an important biological process for the return of fixed
                                 nitrogen to the atmosphere. In the above respiratory elec-
                                 tron transport process electron donors are usually organic
                                 matter or reduced sulfur compounds. Nitrate can also be
                                 reduced to ammonium, but this requires highly  reducing
                                 conditions and organic carbon. The denitrifying  bacteria
                                 use nitrate, nitrite, or nitrous oxide as terminal electron ac-
                                 ceptors for respiratory growth in the presence of  gaseous
                                 oxygen and under micro-aerobic and anaerobic conditions.
                                 Denitrification involves four consecutive steps in which ni-
                                 trate is reduced to dinitrogen gas by the metalloenzymes,
                                 nitrate reductase, nitrite reductase, nitric oxide reductase,
                                 and nitrous oxide reductase:
                                  NCy
      NO2-
        NO
N2O
(Eq. 8.2)
                                 Many denitrifying organisms do not contain or express all
                                 of the four reductases needed for complete denitrification
                                 so that the interaction of different organisms is needed to
                                 reduce nitrate completely into dinitrogen gas.

                                 Anaerobic conditions in the immediate cell environment
                                 are essential for denitrification to occur (Table 8.2). This
                                 translates to a measured  Eh potential  of 300 to 350 mV
                                 (Reddy and Patrick, 1984). Denitrification is not limited to
                                 water-logged soils, it occurs in  moist, fine-textured soils as
                                 well as in localized micro-anoxic zones within the overall
                                 aerobic soil matrix.  The optimum temperature range for
                                 denitrification is from 20 to 35 C and the optimum pH is
                                 neutral to slightly alkaline, but denitrification will occur over
                                 at least pH 5 to 9 (Focht and Verstraete, 1977). Denitrifica-
                                 tion can be described by a zero-order kinetic reaction at
                                 high concentrations of nitrate, and by a first-order reaction
                                 at low concentrations (Focht and Verstraete, 1977).

                                 There are two physiologically distinct classes of denitrify-
                                 ing bacteria: heterotrophic  denitrifiers, which use organic
Table 8.2  Natural attenuation and mobilization pathways for nitrate.
     Attenuation
      Processes
      Mobilization Processes
           Characterization Approach
 Biotic (microbial) or
 abiotic transforma-
 tion of nitrate to other
 nitrogen species.
Cessation of microbial processes via
changes in nutrient supply within the
ground-water flow path; reduction
in the mass of reduced Fe-bearing
minerals within the aquifer due to
changes in redox conditions or supply
of soluble  iron.
Evaluation of N speciation in ground water. Deter-
mination of water chemistry, microbial populations,
and/or sediment mineralogy linked to nitrate trans-
formation along with spatial and temporal variability
of required chemical/microbial components relative
to nitrate transport pathway(s).
                                                     90

-------
compounds as electron donors and carbon sources (Equa-
tion 8.3 and 8.4); and chemolithoautotrophic (or autotrophic)
denitrifiers, which use inorganic compounds (e.g., minerals
containing reduced sulfur and iron) as electron donors and
CO2 as a carbon source (Equations 8.5 and 8.6):
Heterotrophic denitrification:
 4 NO3- +  5CH2O (organic matter)
                                  -
8 NO,- +  5CH3COO-(acetate)  +  3H+
     3
     4I\L
              10HCCy
         4H2O
Autotrophic denitrification:
  14NO3-  + 5FeS2 (pyrite) +  4H+
  7N
          5Fe2+
10SO42-
+ 2H2O
      NO
              5Fe2+
SFeOOH (goethite)  +
     7H20
     0.5 N
                                  9H+
(Eq. 8.3)


(Eq. 8.4)



(Eq. 8.5)


(Eq. 8.6)
Because the rate and extent of denitrification are dependent
on available carbon supply (Reddy and  Patrick, 1984), it
is often assumed that denitrification is not important as a
nitrate-removal mechanism below the root zone. Denitri-
fication is  usually not  significant in the subsurface of low
organic matter and of  sandy soils.
Nitrate was found in  the ground water beneath poorly
drained soils of the North Carolina Coastal Plain and de-
nitrification was proposed as a nitrogen loss mechanism in
the saturated subsoil (Gilliam etal., 1974). Based on nitrate
concentration data, nitrogen mass balances, soluble organic
carbon, and redox potential measurements, Gambrell et al.
(1975) concluded that denitrification could readily occur in
the subsurface (1  to 2.5  m)  of poorly drained subsoils of
the North  Carolina Coastal  Plain. They  attributed this to
the high amounts of dissolved organic carbon (the subsoil
was  overlain by an organic soil), low dissolved  oxygen,
and low Eh (350 mV at 1 m).The nitrate concentration and
nitrate-to-chloride concentration  ratios declined sharply
with  depth. The North Carolina case is relevant to many
other riparian and wetland areas that are believed to  be
significant in denitrification losses at the watershed level.
Under certain flow and redox conditions, riparian zones
along waterways have demonstrated a natural proficiency
to  intercept and denitrify nitrate in shallow ground water
(Gilliam, 1991; Schipper et al., 1991).
The rate and mechanism of nitrate removal along and be-
tween ground water flow paths were investigated using a
series of well nests screened in an unconfined sand and
gravel aquifer in southwestern British  Columbia,  Canada
(Tesoriero et al., 2000). Intensive agricultural activity in this
area has resulted in nitrate concentrations in ground water
often exceeding drinking water standards. Both the extent
and rate of denitrification varied depending on the ground
water flow path. While little or no denitrification occurred
in  much of the upland portions of  the aquifer, a gradual
redox gradient  was observed as aerobic upland ground
water moved deeper into the aquifer. In contrast, a sharp
shallow redox gradient was observed adjacent to a third-
order stream as aerobic ground water  entered reduced
sediments. An essentially complete loss of nitrate concur-
rent with increases in excess dissolved  N2 gas provided
evidence  that denitrification occurred as ground  water
entered this zone. Low denitrification rates were observed
along the deep flow path (< 0.04 umol cm-3 aquifer yr1).
The estimated denitrification rate for the redoxcline  in this
aquifer ranged from 0.3 to 0.8 umol crrv3 aquifer yr1, and
an in-situ experiment  conducted adjacent to the stream
suggested potential denitrification rates may be as high as
42 umol cm-3 aquifer yr1. Electron and mass balance cal-
culations suggested that iron sulfide (e.g., pyrite) deposits
and to a lesser degree  organic matter were electron donors
for denitrification.

Shallow anaerobic ground water not immediately adjacent
to streams exhibits variable but generally low rates  of de-
nitrification. Natural gradient tracer tests  yielded potential
denitrification rates ranging from  0.44 to 1.1  umol cm-3
aquifer yr1 for a shallow sand and gravel aquifer on Cape
Cod, Massachusetts (Smith et al.,  1996). Greater denitri-
fication  rates (24 jj,mol cm"3 aquifer yr1)  were estimated
from an in-situ experiment conducted 3 m below the land
surface  in a shallow aquifer in Ontario, Canada (Trudell et
al., 1986).

Pyrite-bearing aquifers represent important hydrological
compartments due to their capacity to eliminate nitrate. In
the absence of molecular oxygen, the nitrate is reduced
coupled with the microbial oxidation of sulfur in the pyrite
by the bacterium Thiobacillus denitrificans (Pauwels et al.,
1998, Grimaldi et  al.,  2004). A field tracer test using Br
coupled with  nitrate injection revealed that denitrification
rates  in a pyrite-bearing schist aquifer varied depending
on the medium permeability, with slower the denitrification
rate observed in regions of faster ground-water flow velocity
(Pauwels et al., 1998). Denitrification was shown to be ac-
companied by the precipitation of sulfate and iron-bearing
minerals, probably jarosite (KFe3(SO4)2(OH)6).

Under anaerobic conditions, the presence of both nitrate
and ammonium may facilitate establishment of anaerobic
ammonium-oxidizing  (anammox)  bacteria that  directly
oxidize ammonium to  dinitrogen gas with nitrate (Mulder
et al., 1995)  and  nitrite (Dalsgaard et al.,  2003)  as the
electron acceptor:
5NH
                    3NCy
                                                        4N
                                          NH
                                                                    NO2-
                                                 _    9H20  -

                                                 N2  +2H2O
                                                                                          2H+
                                             (Eq. 8.7)

        iA   i  i nvxp   /  i "p  ' ** |pv-'             v M' o.O)

It remains to be seen if ammonium accumulation is an issue
of concern where nitrate is a major remediation target.

Chemical reduction of nitrate has been shown to occur in
a mixture of dissolved Fe2+ and Fe(ll)-containing silicates
(arfvedsonite and augite), but only after an oxyhydroxide
precipitate  had formed on the silicate surface  (Postma,
1990).  Similarly, nitrite has been  shown to be reduced
to N2O by adsorbed Fe(ll) on the surface of lepidocrocite
(y-FeOOH) at pH > 7 (S0rensen and Thorling, 1991). The
Fe(ll) associated with a reactive complex formed during
Fe2+ binding to lepidocrocite, and not the ionic Fe2+, seemed
responsible for the reduction of NO2- similar to the follow-
                                                     91

-------
ing reaction, where the first term in the reaction signifies
ferrous iron adsorbed to lepidocrocite.
      6Fe2+
                2 NO- + 5HO -
            3  4(s)
                                             (Eq.8.9)

In this reaction, it is presumed that the ferrous iron-lepido-
crocite surface complex is converted to surface precipitate
similar in nature to the mineral  magnetite. The catalytic
effect of Fe(lll) oxyhydroxide may stimulate Fe(ll)-depen-
dent formation of N2O from NO2~ (chemodenitrification) in
sediments and subsoils.

In sulfate green rust [Fe(ll)4Fe(lll)2(OH)12SO4. nH2O] sus-
pensions,  nitrate  is stoichiometrically reduced  to NH4+,
and magnetite is the sole Fe-containing product (Hansen
etal., 1996).
4Fe(ll)4Fe(lll)2(OH)
   NH4+ + 8Fe304
12SO4 + NO3- + 6OH-
  4SO42- + 24H2O
                                           (Eq. 8.10)
Recent  thermodynamic and spectroscopic studies  give
direct evidence for the existence of green rusts in soils
(Hansen et al., 1994;Trolard et al., 1997). Since microbial
processes may be expected only to dominate in environ-
ments rich in organic carbon, the green rust-facilitated NH4+
formation may be an important pathway for nitrate removal
and nitrogen conservation in certain anoxic subsoils and
sediments poor in organic matter.

Site Characterization

Groundwater geochemistry provides information about the
source, transformation, and attenuation of nitrate in ground
water.  In  addition, specific soil properties are important
parameters determining the potential for natural attenuation
of nitrate. For example,  nitrate concentrations at the 0.8 to
8 m depth at 15 sites within a beef feedlot were found to
decrease as clay content increased (Lund et al., 1974).
Other characteristics, such as  a clay layer that severely
restricted  water movement,  also decreased soil nitrate
concentrations.  It was probable that considerable soluble
carbon was leached from the feedlot floor to provide energy
for denitrification. The presence of ample organic carbon in
the soil leads to active denitrification. Studies on the fate of
nitrate below organic waste disposal sites (food processing
waste lagoons, manure lagoons, feed yard, sewage sludge,
and effluent septic tank  drainfields) have shown significant
denitrification in the vadose zone (Keeney, 1981).

Measurement of  Inorganic Nitrogen Species

It  is recommended  that site characterization  include
measurement of the various inorganic nitrogen species
- largely nitrate, nitrite,  and ammonia in ground water. At
circum-neutral pH typical of most ground water, ammonia
is  present  predominantly as the ammonium ion (NH4+).
Nitrogen is  a redox  sensitive element  that is involved
in  numerous chemical  and  biological processes.   It  is
anticipated that the  fate and  transformation of nitrate,
nitrite, and ammonia  in  the subsurface would be reflected
by the changes of their concentrations in the  aqueous
phase of  the subsurface environment with time.   It  is
therefore important to monitor their concentrations as an
integral part of the overall evaluation of natural attenuation
at a site. A summary of available laboratory methods for
determination of the various inorganic species of nitrogen
in ground water is provided in Table A.1 in USEPA (2002),
which  is available at http://www.epa.gov/ada/download/
reports/epa_600_r02_002. pdf.

Although field colorimetric methods are available to quan-
tify nitrate, nitrite, and ammonia, field determination using
these methods is not generally needed as long as labora-
tory measurements can be made within specified holding
times.  Samples collected for nitrogen speciation are to
be preserved by keeping cold (4C) and measured within
specified holding times,  generally < 48 h. Depending on
the method of analysis, longer holding times are possible
by acidifying samples with sulfuric acid and keeping them
cold. Prior to analysis, acid-preserved samples are to be
brought back to room temperature and neutralized by add-
ing base.

Use of Nitrogen Isotope Ratios for
       Identifying Sources and  Transformation
       Processes of Nitrate

There are two stable isotopes of nitrogen: 14N (with a natural
abundance of 99.6337 atom percentage) and 15N (0.3663)
(Junk and Svec, 1958). Enrichment of  15N is expressed on
the per mille (%<>) basis, and is calculated as:

Si Woo = (atom%i5Nsample - atom%^Nstandard) x
        1000/(atom%i5Nstandard)              (Eq.8.11)

The standard  is usually atmospheric nitrogen. Manufac-
tured fertilizers and nitrate in precipitation tend to have 15N
concentrations close to  natural  abundance (815N%0 = 0),
and soil organic matter nitrogen and hog manure organic
nitrogen show greater 815N%0 values  (Kellman, 2005). In
bacterially mediated  processes  the reaction of the  light
isotopic species is kinetically favored. As a result, the 15N/14N
ratio of the instantaneously-formed product  is lower than
the 15N/14N ratio of  the  remaining, unreacted  substrate.
Ammonia volatilization can also  lead to 15N isotope en-
richment in the remaining substrate. This "kinetic isotope
fractionation" is defined in terms of the isotope enrichment
factor,  Eproduct-substrate (p-s)>
                                                                 815N
                                                                     product
                         S15N
                                                                                (Eq.8.12)
                                                       For bacterial processes, e  is negative or zero (Heaton
                                                       et al., 2005). If a process can be approximated as a first
                                                       order reaction, changes in the 515N value of the product or
                                                       substrate of the reaction can be calculated using "Rayleigh"
                                                       equations (Heaton et al., 2005).

                                                       For the substrate,
                                                            S8.t  =  8s,o   +   ep-s-ln(f)>             (Eq.8.13)
                                                       and for the total accumulated product,
                                                          Spt   =  SSO  -   ep.s.ln(f)[f/(1-f)]          (Eq.8.14)

                                                       Where  8S,, 8p, are the 815N values of the remaining sub-
                                                       strate and accumulated product at time t; 8s, 0 is the initial
                                                     92

-------
515N value of the substrate; f is the fraction of substrate
remaining; and ep s is the isotope enrichment factor for the
process.

Since the early 1990s, a few published studies have focused
on the transport and fate of nitrate in ground water using
dual isotope analysis (e.g., 815N and 818O in dissolved NO3")
in combination with other lines of evidence to document
in-situ microbial denitrification (Aravena and Robertson,
1998; Beller et al., 2004). Analytical procedures have been
developed for determination of 815N via catalytic reduction of
nitrate to N2 after anion exchange separation and precipita-
tion as AgNOg, and for determination of 818O via combus-
tion of nitrate to CO2 using excess graphite (Aravena and
Robertson, 1998). For determination of isotopic signatures
of both nitrate and nitrite, a quantitative method has been
developed based on bacterial denitrification  (Sigman et
al., 2001; Casciotti et al., 2002). Widory et al. (2005) used
a coupled isotopic approach (815N  and 811B), in addition to
conventional hydrogeological analyses, to trace the origin of
NOg- in ground water. The studied watersheds include both
fractured bedrock and alluvial (subsurface and deep)  hy-
drogeological contexts. The joint use of nitrogen and boron
isotope systematics in each context deciphers the origin of
NOg~ in the groundwater and allows a semi-quantification of
the contributions of the respective pollution sources (mineral
fertilizers,  wastewater, and animal manure).

To investigate the fate of nitrate in a petroleum-contami-
nated aquifer, 15N isotope and acetylene-inhibition methods
in combination with single-well push-pull tests were used
to quantify processes contributing to nitrate consumption
(Schurmann et al., 2003).The processes quantified included
denitrification, assimilatory nitrate  reduction, dissimilatory
nitrate reduction to ammonium, and abiotic nitrate reduc-
tion. Multiple lines of evidence from chemical composition,
stable isotope (13C/12C, 15N/14N, and 34S/32S), and dissolved
gas (N2, Ar, O2, and CH4) composition of ground water at a
landfill site were used to account for all NH4+ loss by com-
bined nitrification and denitrification processes in  a system
where there were abrupt temporal and spatial changes in
redox conditions (Heaton et al., 2005).

There are pitfalls in  nitrogen source identification  us-
ing  the  nitrogen isotope approach. The heterogeneity of
complex watersheds makes source identification difficult;
the  complexity of the  nitrogen cycle and the associated
isotope fractionations can make the results ambiguous or
unreliable; source identification becomes very difficult when
mixing of point and non-point sources occurs; isotopic ratios
sometimes are hard to determine in environmental samples
due to analytical difficulties. It appears that use of 815N to
identify  sources of nitrogen to ground waters is feasible
only in relatively simple systems  (e.g., low rainfall, single
or dual sources, and minimal nitrogen transformations such
as denitrification). Watershed- or ecosystem- level studies
using this technique are semiquantitative at best. In most
cases where  815N can be applied reliably, the sources of
nitrogen are obvious or can readily be estimated by other,
less expensive means (Keeney, 1986). In one case, nitrogen
isotopic data were of little help in determining the relative
importance of cyanide and explosive residues as nitrogen
sources in surface ponds at three Nevada gold mine sites.
The data do, however, provide strong evidence for natural
attenuation of non-cyanide  nitrogen species  by dispersal
of ammonia gas or by dispersal of N2O or N2 gas produced
by denitrification of nitrate (Johnson et al., 2000).

Use of 1H Nuclear Magnetic Resonance for
      Identifying Sources of Nitrate
Dissolved organic matter (DOM) originating from a certain
source usually carries characteristic marks in its molecular
structures that can be recognized by spectroscopic analysis
such as nuclear magnetic resonance (NMR). Sources of
water-born contaminants, such as nitrate, can be identified
by recognition of the characteristics of DOM  entrained in
the water. Lu et al. (2004) analyzed DOM in ground water
collected from a dairy/crop production area (Chino Basin,
CA) using 1H NMR. Results showed that DOM derived from
natural soil organic matter has a characteristic resonance
at a chemical shift region of 4.0-4.3 ppm, while DOM de-
rived from dairy wastes has a characteristic resonance at
a lower chemical shift region of 3.2-3.6 ppm. These signa-
ture resonances were then used to distinguish the origins
of nitrate in the ground water.  It was found that disposal
of dairy wastes on croplands was the primary source of
nitrate contamination in ground water underlying the Chino
Basin dairy area.

Denitrification Enzyme Activity
Denitrification enzyme activity (DBA)  experiments can be
conducted according to Tiedje (1994) using soils/sediments
collected from the aquifer.  It is recommended that these
tests be conducted with caution due to potential disturbance
to intact soil/sediment conditions during retrieval and pro-
cessing.  However,  DBA has been shown  to be strongly
related to annual denitrification rates in temperate zone
soils,  and is useful for comparisons within a site as well
as for the distribution of rates within soil horizons (Groff-
man and Tiedje, 1989; Schnabel et al., 1996; Flite et al.,
2001). The end product of denitrification is N2 gas, thus,
it may be useful to analyze ground water for dissolved N2
gas at locations where  soils/sediments  are  sampled for
DBA experiments. This dissolved  gas is a  good indicator
of denitrification when found in excess of concentrations
expected relative to atmospheric trace gases such as Ar
(Vogel et al., 1981; Bohlke and Denver, 1995).

Estimation of Denitrifying Capacity
Ground-water denitrification capacity  may be  evaluated
through laboratory microcosm studies using aquifer sedi-
ments or through direct in-situ tracer tests in discrete loca-
tions within the aquifer. However, successful extrapolation
of the results from these measurements across the site will
depend on the  degree of sampling frequency to capture
spatial heterogeneity in both horizontal and vertical dimen-
sions of the site with respect to hydrology,  soil character-
istics, and biogeochemical processes. It is  recommended
that both approaches be employed for a given site, since
there is evidence that measured potentials of nitrate removal
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derived from laboratory microcosms may overestimate re-
moval capacities in the field (Kellman, 2004). Korom et al.
(2005) illustrated the use of in-situ mesocosm to determine
denitrification rates  in aquifer sediments.  These authors
incorporated information from tracer tests, 815N data, and
geochemical data to determine that 58% of the denitrifica-
tion was caused by S(-l) in pyrite as the electron donor with
the rest from ferrous iron and organic carbon.

To quantify ground  water denitrification in discrete loca-
tions of riparian aquifers, Addy et al. (2002) modified and
evaluated an in-situ  method based on conservative tracers
and 15N-enriched nitrate. Ground water was "pushed" (i.e.,
injected)  into a  mini-piezometer and then "pulled" (i.e.,
extracted) from the same mini-piezometer after an incuba-
tion period. This push-pull method was applied in replicate
mini-piezometers at two Rhode  Island  riparian sites, one
fresh water and one brackish water. Conservative tracer
pretests were conducted to determine incubation periods,
ranging from 5 to 120 h, to optimize recovery of introduced
plumes. For nitrate push-pull tests, they used two conser-
vative tracers, sulfur hexafluoride and bromide, to provide
insight into plume recovery. The  two conservative  tracers
behaved similarly.The dosing solutions were amended with
15N-enriched nitrate to quantify the mass of denitrification
gases generated during  the incubation  period. The in-situ
push-pull method detected substantial denitrification rates
(mean  of 97 ug N kg-1 d~1) at a depth of 65 cm in a glacial
outwash setting  located within 10 m of the stream. At the
brackish site, high rates of ground water denitrification were
found in marsh  locations (mean  of 123 ug N kg~1 d~1) and
minimal denitrification in soils fringing the marsh (mean of
2 ug N  kg~1 d~1) at a depth of 125 cm. The push-pull method
can provide useful insights into spatial  and temporal pat-
terns of denitrification in  riparian zones. The  method is
robust  and results are not seriously affected by dilution or
degassing from ground water to soil air. In conjunction with
measurements  of ground water flow, this method holds
promise for evaluating the influence of site and manage-
ment factors on the ground-water nitrate removal capacity
of riparian zones (Addy et al., 2002; Kellogg et al., 2005).

Attempts have been made to estimate the denitrifying re-
moval capacity of an entire wetland (MaTtre et al., 2005).
The authors developed  a methodology that  consisted of
the following steps:  1) delineating the structure of aquifer
formation in the riparian wetland; 2) mapping the spatial
variation  of the  thickness of the different soil horizons
constituting the aquifer formation; 3) measuring the DBA
for each type of soil horizon; 4) grouping the  soil horizons
in specific soil-denitrifying classes; 5) mapping the spatial
variation  of the  thickness of  each soil-denitrifying class;
6) mapping the spatial variation of the water table posi-
tion; 7) combining the maps to calculate the volume of the
fraction of each soil-denitrifying  class that interacts with
groundwater; and 8) calculating the denitrifying removal
capacity of the whole site.

Long-term Capacity

The long-term capacity of  nitrate transformation within the
aquifer will depend on the prevailing ground-water chem-
istry and the dynamics of biotic and  abiotic  processes
controlling conversion of nitrate to other nitrogen species.
If there are significant changes in ground-water chemistry
over time, then the potential exists for  reductions in the
aquifer denitrifying capacity relative to nitrate loading from
the contaminant plume.   For situations in which  nitrate
transformation is controlled by biotic (microbial) processes,
this may be due to reductions in degradable organic carbon
(dissolved or associated with aquifer matrix) and nutrients
needed to sustain the microbial community. For situations
where abiotic components control  nitrate transformation,
e.g., reduced Fe-bearing minerals, changes to more oxidiz-
ing conditions or reductions in the supply of iron to sustain
the mass of reactive minerals would likely  result in nitrate
plume expansion. Ultimately, the evolution of biogeochemi-
cal conditions in the subsurface will dictate the success of
the MNA remedy.

Since heterotrophic denitrification reactions require that the
water is anaerobic (Eh < 350 mV) and that  considerable
organic carbon is present at concentrations greater than the
nitrate concentration (Korom, 1992), significant widespread
denitrification in aquifers are unlikely to occur.  Based on
water chemistry data, it  was concluded that  denitrifica-
tion  cannot  be relied upon to decrease elevated  nitrate
concentrations  in the modern, polluted, recharge waters
of the English chalk (limestone) aquifer (Howard,  1985).
On the other hand,  situations exist where denitrification
may be occurring in some shallow aquifers. For example,
Egboka (1984) studied six ground water flow systems in
Ontario, Canada under widely varying land use.  Several
of the ground waters exhibited low dissolved oxygen and
Eh, but there was little consistency in the nitrate to chloride
ratio. Nitrate concentrations at the sites that were reduc-
ing (low Eh) declined with distance from the nitrate source.
Vidon and  Hill  (2004) observed effective nitrate removal
by denitrification in riparian zones with hydric soils as well
as in non-hydric riparian zones,  and they  concluded that
a shallow water table is not always necessary for efficient
nitrate removal  by denitrification.

Chemolithoautotrophic denitrification may occur with pyrite
as the electron donor. This is the proposed mechanism
derived from a recent study based on the low dissolved
organic carbon concentrations (< 1.5 mg L1) that could not
support heterotrophic denitrification, the common occur-
rence of disseminated pyrite in the aquifer, and the trend
of increasing sulfate as ground water flowed from aerobic,
unconfined conditions to anoxic, confined aquifer conditions
(Beller et al., 2004).  In this study, several independent
lines of evidence suggested that microbial denitrification
was naturally attenuating nitrate in a confined, O2-depleted
region of a bedrock aquifer at a Lawrence Livermore Na-
tional Laboratory site. The evidence included the following
observations: (1) both nitrate and dissolved oxygen concen-
trations in ground water decreased dramatically as ground
water flowed from unconfined to confined aquifer conditions,
(2) stable isotope signatures (i.e., 815N and 818O) of ground
water NO3- indicated a trend of isotopic enrichment that was
characteristic of denitrification, and (3) dissolved N2 gas,
the product of denitrification, was highly elevated in nitrate-
                                                      94

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depleted ground water in the confined region of the aquifer.
At this site, long-term nitrate concentrations were relatively
high and constant in recharge-area monitoring wells (typi-
cally 70-100 mg NO3' L1) and relatively low and constant
in the downgradient confined region (typically < 0.1 - 3 mg
NO3~ L'1), suggesting a balance between rates of  nitrate
loading  and removal by denitrification.  Autotrophic deni-
trification with pyrite (or other reduced Fe-S minerals) has
been invoked to explain geochemical results of other field
studies where low dissolved organic carbon concentrations
could not explain apparent nitrate degradation in ground
water (Kolle et al., 1985; Bottcher et al., 1990; Postma et al.,
1991; Robertson et al., 1996; Pauwels et al., 1998; Aravena
and Robertson, 1998; Tesoriero et al., 2000).

Tiered Analysis

Determination of the viability of nitrate remediation in ground
water via monitored natural attenuation will depend upon
proper assessment of contaminant loading to the aquifer
and  prevailing  geochemistry and mineralogy within the
contaminant plume and the down gradient zone prior to
the point(s) of compliance.  It  is essential to understand
the cause-and-effect relationship between loss of nitrate in
ground water and the mechanisms responsible for its loss.
The following tiered analysis structure for site characteriza-
tion provides  an approach to evaluate candidate sites and
define the potential limitations of MNA as part of a remedy
for ground-water cleanup.

Tier I. Site characterization under Tier  I will involve dem-
onstration that  the plume is static or shrinking,  has not
reached compliance boundaries, and  does not impact
existing  water supplies.  Once  this is established through
ground-water characterization, evidence is collected to
demonstrate  nitrate reduction  within the plume.   Rapid
movement of contaminants along preferred flow  paths in
the unsaturated and  saturated zones can be induced by
hydrologic events such as heavy rains (e.g., McCarthy et
al., 1998; Camobreco et al., 1996).  It will be important to
determine that such hydrogeologic features do not result in
contaminants bypassing zones where natural attenuation is
occurring.  An observed decrease in nitrate concentration
in space and time that is attributable to a mass-removal
process is anticipated if natural attenuation processes are
active throughout the plume. Conditions that would support
this initial screening would include evidence of reducing
conditions (i.e., low EH) and abundant electron donors (e.g.,
elevated total organic matter or the presence of degradable
organic  co-contaminants). Identification of nitrate removal
along ground-water flow paths  provides justification for
proceeding to Tier II characterization efforts.

Tier II. Under Tier II, the apparent rate and mechanism(s)
of attenuation are determined.  Estimates of a site attenu-
ation  rate(s) can be determined via  several approaches:
1) assessment of nitrate disappearance across well tran-
sects along the ground-water flow path(s), 2) assessment
of in-situ rates of nitrate reduction  through  analysis of
changes in the nitrogen stabile isotope fractionation along
the ground-water flow path, and 3) ex-situ assessment of
nitrate reduction rate(s) through the use of microcosm stud-
ies. In addition, time-series data may be collected at one or
more monitoring points within the plume. This information
will allow assessment of the relative timescale(s) for nitrate
reduction relative to ground-water seepage velocities and
determination of whether remediation objectives can be met
within the required regulatory time frame.  In addition, the
mechanism(s) for attenuation need to be identified under
this stage of site characterization (e.g., See site character-
ization approach in Postma et al., 1991.). This effort may re-
quire identification of the microbial agents involved in nitrate
reduction and their distribution throughout the plume (see
Site Characterization section in this chapter). Alternatively,
sites at which an abiotic process dominates nitrate reduction
may require characterization of aquifer solids to determine
the presence of candidate mineral phases such as those
discussed previously (see Equations 8.10  and 8.11).  If a
link is established between the apparent  disappearance
of nitrate and observed microbial or chemical processes
active within the plume, then this provides justification for
proceeding to Tier III characterization efforts.

Tier III. Once the nitrate reduction mechanism(s) have been
identified for the site, the subsequent characterization effort
under Tier III will involve determination of the capacity of
the aquifer to sustain continued biological reduction.  The
impact of potential hydrologic changes, such as a shift in
flow  direction caused by the onset of pumping at nearby
sites, needs to be determined. Since the active microbial
community will reside within the boundaries of the nitrate
plume, potential alterations in the predominant flow path
may deliver  nitrate to regions of the aquifer in which there
is a  reduced (or inactive)  population of nitrate-reducing
bacteria  (NRB) to sustain reduction.  A well-constrained
ground-water flow model of the site  relative to observed
spatial distributions of nitrate and NRB density will assist
in assessing the potential impact of such changes.  For
sites  in which organic co-contaminants are  identified as
the predominant electron donor for nitrate reduction, it is
recommended that an evaluation of the long-term supply
of the electron donor be assessed. If engineered remedies
for the treatment of co-contaminants are active or planned
(e.g., source zone treatment or removal  of co-contami-
nants), then it is expected that the impact these remedies
have on the potential supply of electron donor will influence
the sustainability of  nitrate reduction.  This  may  require
evaluation of whether 1) the total  organic carbon content
and characteristics of treated ground water are sufficient to
maintain NRB activity or 2)  the available mass of reactive
minerals within aquifer sediments is sufficient to sustain
chemical reduction.  Analysis of the capacity for sustained
biological reduction could be conducted through laboratory
microcosm studies, whereas assessment of the long-term
capacity for  chemical reduction may be achieved through
assessment of  the mass of reactive minerals within the
plume flow path and their stability relative to anticipated
changes in ground-water chemistry.

For sites in which uncontaminated portions of the aquifer
are aerobic (i.e., oxygen is present)  or  have naturally
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elevated concentrations of competing electron acceptors
such as perchlorate,  it is anticipated that evolution of the
plume geochemistry may evolve to a condition that will not
sustain microbial nitrate reduction. It is recommended that
the potential impact of anticipated changes in ground-water
chemistry on the rate or sustainability of nitrate reduction
be assessed through microcosm studies.  Ultimately, the
ability to forecast the potential impacts of changes in aquifer
chemistry or ground-water flow can be improved through
the development of a reaction-transport model that includes
site-specific parameterization of the microbial or chemical
reduction  process and  site hydrogeology.  If site-specific
tests demonstrate sufficient capacity within the aquifer to
sustain nitrate reduction, then the site characterization effort
can progress to Tier IV. For cases where aquifer capacity
is insufficient for plume reduction to required levels, then a
determination of the benefits of contaminant source reduc-
tion or removal may be necessary.

Tier IV.  Finally, under Tier IV a monitoring plan is estab-
lished along with contingency plans in the event of  MNA
failure.   It is recommended  that  the monitoring plan be
designed to establish both continued plume stability and to
identify changes in ground-water  chemistry that may lead
to a decrease in the rates or capacity for nitrate reduction.
The specific chemical  parameters to  be  monitored will
include those identified under Tier III that may halt or slow
down microbial nitrate reduction within the plume. For ex-
ample,  solution phase parameters that could alter nitrate
reduction include the concentrations of dissolved oxygen or
competing electron acceptors such as perchlorate.  Simi-
larly, a decrease in the concentration of electron donors
such as natural organic matter, organic co-contaminants,
or  reactive minerals  could slow or halt nitrate reduction
via biological or abiotic processes.  Changes in these pa-
rameters may occur  prior to observed changes in nitrate
concentrations and, thus, serve as monitoring triggers for
potential MNA failure. In this instance, a contingency plan
can be implemented  that incorporates strategies to arrest
possible plume expansion beyond compliance boundaries.
Possible strategies to prevent plume expansion include
pump and treat operations, installation of reactive barriers
to enhance reduction capacity perpendicular to the direction
of plume advance, or enhancement of natural attenuation
processes within the aquifer through the injection of soluble
reactive components  (see Remedial Technologies section
in this chapter).

References

Addy, K., D.Q. Kellogg, A.J. Gold, P.M. Groffman, G. Ferendo,
    and C. Sawyer. In situ push-pull method to determine
    ground water denitrification in riparian zones. Journal
    of Environmental Quality 31:1017-1024 (2002).
Alowitz, M.J. and M.M. Scherer.  Kinetics of nitrate, nitrite,
    and Cr(VI) reduction by iron metal. Environmental Sci-
    ence and Technology 36: 299-306 (2002).
Aravena,  R. and W.D. Robertson. Use of multiple isotope
    tracers to evaluate denitrification in ground water: Study
    of nitrate from a large-flux septic system  plume. Ground
    Water 36: 975-982 (1998).
Seller,  H.R., V Madrid, G.B. Hudson, W.W.  McNab, and
    T. Carlsen,  Biogeochemistry and natural attenuation
    of nitrate in groundwater at an explosives test facility.
    Applied Geochemistry 19: 1483-1494 (2004).
Bohlke, J.K. and J.M. Denver. Combined use of groundwater
    dating, chemical, and isotopic analyses to resolve the
    history and fate of nitrate contamination in two agri-
    cultural watersheds, Atlantic coastal plain, Maryland.
    Water Resources Research 31: 2319-2339 (1995).
Bottcher, J., O. Strebel, S. Voerkelius, and H.L. Schmidt.
    Using isotope fractionation of nitrate-nitrogen and ni-
    trate-oxygen for evaluation of microbial denitrification
    in a sandy aquifer. Journal of Hydrology 114:413-424
    (1990).
Boyce, J.S., J. Muir,  A.P. Edwards, E.G. Seim, and  R.A.
    Olson.   Geological nitrogen  in Pleistocene  loess of
    Nebraska. Journal of Environmental Quality 5: 93-96
    (1976).
Burton, S.A.Q. and I.A. Watson-Craik. Ammonia and nitro-
    gen fluxes in landfill sites: Applicability to sustainable
    landfilling. Waste Management  Research 16: 41-53
    (1998).
Buss, S.R., A.W. Herbert, P. Morgan, S.F Thornton, and
    J.W.N. Smith.  A review of ammonium  attenuation in
    soil and groundwater. Quarterly Journal of Engineering
    Geology and Hydrogeology 37: 347-359 (2004).
Camobreco, V.J., B.K. Richards, T.S. Steenhuis, J.H. Peverly,
    and M.B. McBride. Movement of heavy metals through
    undisturbed and  homogenized soil columns. So/7 Sci-
    ence 161: 740-750 (1996).
Casciotti, K.L., D.M. Sigman, M.G. Hastings, J.K. Bohlke,
    and A. Hilkert. Measurement of the oxygen isotopic
    composition of nitrate in seawater and freshwater us-
    ing the denitrifier method.  Analytical Chemistry 74:
    4905-4912 (2002).
Centers for Disease Control and Prevention.  Spontane-
    ous abortions possibly  related  to  the ingestion of
    nitrate-contaminated water-LaGrange County, Indiana
    1991 -1994. Morbidity and Mortality Weekly Report 45:
    569-572 (1996).
Ceplecha, Z.L., R.M. Waskom.TA. Bauder, J.L. Sharkoff,
    and R. Khosla. Vulnerability assessments of Colorado
    ground water to nitrate contamination. Water, Air, and
    Soil Pollution 159: 373-394 (2004).
Chen,  S.S., H.D. Hsu, and C.W.  Li. A new  method to
    produce nanoscale iron for nitrate removal. Journal of
    Nanoparticle Research 6: 639-647 (2004).
Chen, Y.M., C.W. Li, and S.S. Chen.  Fluidized zero valent
    iron bed reactor for nitrate removal. Chemosphere 59:
    753-759 (2005).
Chew, C.F and T.C. Zhang. In situ remediation of nitrate-
    contaminated ground water by electrokinetics/iron wall
    processes. Water Science and Technology 38:135-142
    (1998).
Chew, C.F. and T.C. Zhang. Abiotic degradation of nitrates
    using zero-valent iron and electrokinetic processes. En-
    vironmental Engineering Science 16: 389-401 (1999).
                                                     96

-------
Dalsgaard.T., D.E.Canfield, J.Petersen, B.Thamdrup, and
    J. Acuna-Gonzalez.  N2 production by the anammox
    reaction in the anoxic water column of  Golfo Dulce,
    Costa Rica. Nature 422: 606-608 (2003).
Dorsheimer, W.T.,  C.B.  Drewry, DP. Fritsch,  and  D.E.
    Williams. Removing nitrate from groundwater. Water
    Engineering and Management 144: 20-24 (1997).
Egboka, B.C.E. Nitrate contamination of  shallow ground-
    waters in Ontario,  Canada. Science of the Total Envi-
    ronment 35: 53-70 C\ 984).
Flite III, O.R,  R.D.  Shannon,  R.R. Schnabel,  and R.R.
    Parizek. Nitrate removal in a riparian wetland of the
    Appalachian Valley and Ridge physiographic province.
    Journal of Environmental Quality 30: 254-261 (2001).
Focht,  D.D. and W. Verstraete.  Biochemical ecology of
    nitrification and  denitrification. Advances in Microbial
    Ecology 1: 135-214(1977).
Gambrell,  R.R, J.W. Gilliam, and S.B. Weed. Denitrifica-
    tion in sub-soils of the North Carolina costal plain as
    affected by soil drainage.  Journal of Environmental
    Quality 4:311-316(1975).
Gilliam, J.W., R.B. Daniels, and J. F Lutz. Nitrogen content
    of shallow ground water in the North Carolina coastal
    plain.  Journal of Environmental Quality 3: 147-151
    (1974).
Gilliam, J.W., R.W.  Skaggs,  and S.B. Weed.   Drainage
    control to diminish nitrate loss from agricultural fields.
    Journal of Environmental Quality 8: 137-142 (1979).
Gilliam, R.W.  Nitrate  contamination of ground water  in
    southern Ontario and the evidence for denitrification.
    In Nitrate contamination: Exposure, consequence, and
    control. NATO ASI Series G: Ecological  Sciences 30,
    I. Bogardi and  R.D. Kuzelka (Eds.), Springer-Verlag,
    Berlin, pp. 181-198(1991).
Grimaldi, C., V.Viaud, F. Massa, L.Carteaux, S. Derosch, A.
    Regeard, Y. Fauvel, N. Gilliet, and F. Rouault. Stream
    nitrate variations  explained by  ground water  head
    fluctuations in a pyrite-bearing aquifer. Journal of En-
    vironmental Quality 33: 994-1001 (2004).
Groffman,  P.M. and J.M.Tiedje. Denitrification in north tem-
    perate forest soils: Relationships between  denitrification
    and environmental factors at the landscape scale. So/7
    Biology and Biochemistry 21: 621 -626 (1989).
Gu, B., D.B.Watson, L. Wu, D.H. Phillips, D.C. White, and
    J. Zhou. Microbiological characteristics in a zero-valent
    iron reactive barrier. Journal of Environmental Monitor-
    ing and Assessment 77: 293-309 (2002).
Hamon, M. and E. Fustec. Laboratory and  field study of
    an in situ groundwater denitrification reactor. Research
    Journal of Water Pollution Control Federation 63: 942-
    949 (1991).
Hansen, H.C.B., O.K. Borggaard, and J. S0rensen. Evalu-
    ation of  the free energy of formation of Fe(ll)-Fe(lll)
    hydroxide-sulphate (green  rust)  and  its reduction of
    nitrite. Geochimica et Cosmochimica Acta  58: 2599-
    2608(1994).
Hansen, H.C.B., C.B. Koch, H. Nancke-Krogh, O.K. Borg-
    gaard, and J. S0rensen.  Abiotic nitrate reduction to
    ammonium: Key role of green rust.  Environmental
    Science and Technology 30: 2053-2056 (1996).
Heaton, T.H.E., J.K. Trick, and G.M. Williams.  Isotope and
    dissolved  gas evidence for  nitrogen attenuation in
    landfill leachate dispersing into a chalk aquifer. Applied
    Geochemistry 20: 933-945 (2005).
Hiscock,  K.M., J.W. Lloyd,  and  D.N. Lerner.  Review of
    natural and  artificial denitrification  of groundwater.
    Water Research 25: 1099-1111 (1991).
Holloway, J.M., R.A. Dahlgren, B. Hansen, and W.H. Casey.
    Contribution of bedrock nitrogen to high nitrate concen-
    trations in stream water. Nature 395:  785-788 (1998).
Howard, K.W.F Denitrification in a major limestone aquifer.
    Journal of Hydrology 76: 265-280 (1985).
ITRC (InterstateTechnology & Regulatory  Council). Emerg-
    ing Technologies for Enhanced In Situ Biodenitrifica-
    tion (EISBD) of Nitrate-Contaminated Ground Water.
    Washington, D.C.: Interstate Technology & Regulatory
    Council, Enhanced In S/ft/Biodenitrification WorkTeam,
    69 pp. (2000).
Johnson, C.A., D.J. Grimes, and R.O. Rye.  Fate of  pro-
    cess solution cyanide and nitrate at three Nevada  gold
    mines inferred from stable carbon and nitrogen isotope
    measurements. Transactions of the Institute of Mining
    and Metallurgy, Section C:  Mineral  Processing  and
    Extractive  Metallurgy 109: C68-C78 (2000).
Jordan, T.E., D.L. Correll, and D.E. Weller. Nutrient intercep-
    tion by a riparian forest receiving inputs from adjacent
    cropland. Journal of Environmental Quality 22:467-473
    (1993).
Junk, G. and H.V. Svec. The  absolute abundance of the
    nitrogen isotopes in the atmosphere and compressed
    gas from various sources. Geochimica et Cosmochi-
    mica Acta  14: 234-243 (1958).
Keeney, D.R. Soil nitrogen chemistry and  biochemistry. In
    Modeling Wastewater Renovation: Land Treatment. I.
    K.  Iskandar (Ed.), John Wiley & Sons, New York, NY
    (1981).
Keeney, D.R.  Sources of nitrate  to ground water. CRC
    Critical Reviews in Environmental Control 16:257-304
    (1986).
Kellman, L. Nitrate removal in a first-order stream: Recon-
    ciling laboratory and field measurements. Biogeochem-
    istry 71: 89-105 (2004).
Kellman,  L.M.  A study  of tile drain nitrate-515N values as
    a tool for assessing nitrate sources in an agricultural
    region. Nutrient Cycling in Agroecosystems 71: 131-
    137(2005).
Kellogg, D.Q., A.J. Gold, P.M. Groffman, K. Addy, M.H. Stolt,
    and G. Blazejewski. In situ  ground water denitrifica-
    tion in stratified, permeable soils underlying riparian
    wetlands. Journal of Environmental  Quality 34: 524-
    533 (2005).
                                                    97

-------
Kolle, W., O. Strebel, and J. Bottcher. Formation of sulfate
    by microbial denitrification in a reducing aquifer. Water
    Supply 3: 35-40 (1985).
Korom, S.  Natural denitrification in the saturated zone:
    A review. Water Resources Research 28: 1657-1668
    (1992).
Korom, S.F., A.J. Schlag, W.M. Schuh,  and A.K. Schlag.
    In situ  mesocosms: Denitrification  in the Elk  Valley
    aquifer.  Ground Water Monitoring and Remediation
    25: 79-89 (2005).
Lu, J., A.C. Chang, and L. Wu. Distinguishing sources of
    groundwater nitrate by 1H  NMR of dissolved organic
    matter. Environmental Pollution  132: 365-374 (2004).
Lund, L.J., D.E. Adriano,  and  P.P. Pratt. Nitrate concen-
    trations in deep soil profiles as related to soil  profile
    characteristics. Journal of Environmental Quality 3:
    78-82 (1974).
MaTtre V., A.C. Cosandey, A.  Parriaux, and C. Guenat. A
    methodology to estimate  the denitrifying capacity of a
    riparian wetland. Journal of Environmental Quality 34:
    707-716(2005).
McCarthy, J.F., K.R. Czerwinski, W.E. Sanford, P.M. Jardine,
    and J.D. Marsh. Mobilization of transuranic radionu-
    clides from disposal trenches by natural organic matter.
    Journal of Contaminant Hydrology 30: 49-77 (1998).
McKeon, C.A., FL. Jordan, E.P Glenn,  W.J. Waugh, and
    S.G. Nelson. Rapid nitrate loss from a contaminated
    desert soil. Journal of Arid Environments 61: 119-136
    (2005).
Mishra, D. and J. Farrell. Understanding nitrate reactions
    with zerovalent iron using Tafel analysis and electro-
    chemical impedance  spectroscopy.  Environmental
    Science and Technology  39: 645-650 (2005).
Mulder,  A.,  A.A. Vandegraaf,  L.A.  Robertson,  and  J.G.
    Kuenen. Anaerobic ammonium oxidation discovered in
    a denitrifying fluidized bed reactor. FEMS Microbiology
    and Ecology 16: 177-183 (1995).
National Research Council. Alternatives for Ground Water
    Cleanup. National  Academy Press, Washington, DC.,
    (1994).
Nolan, B.T. Relating nitrogen  sources and aquifer suscep-
    tibility to nitrate in shallow ground waters of the  United
    States. Ground Water 39: 290-299 (2001).
Nolan,  B.T. and J.D.  Stoner. Nutrients  in ground waters
    of the conterminous United  States, 1992-1995. En-
    vironmental Science and Technology 34: 1156-1165
    (2000).
North, N.N., S.L. Dollhopf, L. Petrie, J.D. Istok, D.L. Balkwill,
    and J.E. Kostka. Change  in bacterial community struc-
    ture during in situ biostimulation of subsurface sediment
    contaminated with uranium and nitrate. Applied and
    Environmental Microbiology 70: 4911 -4920 (2004).
Omernik, J.M. The Influence of Land Use on Stream Nutri-
    ent Levels, EPA-600/3-76-014. Environmental Research
    Laboratory, Corvallis, OR: U.S. Environmental Protec-
    tion Agency, (1976).
Pauwels, H., W. Kloppmann, J.C. Foucher, A. Martelat,
    and V. Fritsche. Field tracer test for denitrification in
    a pyrite-bearing schist aquifer. Applied Geochemistry
    13:767-778(1998).
Postma, D.  Kinetics of nitrate reduction by detrital Fe(ll)
    silicates. Geochimica Cosmochimica Acta 54:903-908
    (1990).
Postma, D., C. Boesen, H. Kristiansen, and FLarsen. Nitrate
    reduction in an unconfined sandy aquifer: Water chem-
    istry, reduction processes, and geochemical modeling.
    Water Resources Research 27: 2027-2045 (1991).
Puckett, L.J.  Identifying the major sources of nutrient water
    pollution. Environmental Science and Technology 29:
    408A-414A(1995).
Reddy, K.R. and W.H. Patrick. Nitrogen transformation and
    loss in flooded soils and sediments. CRC Critical Re-
    views in Environmental  Control 13: 273-309 (1984).
Robertson, W.D. and J.A. Cherry. In situ denitrification of
    septic-system nitrate using reactive porous media bar-
    riers: Field trials. Ground Water 33: 99-111 (1995).
Robertson, W.D., B.M. Russell, and J.A. Cherry. Attenuation
    of nitrate in aquitard sediments  of southern Ontario.
    Journal of Hydrology 180: 267-281 (1996).
Robertson, W.D., D.W. Blowes, C.J. Ptacek, and J.A. Cherry.
    Long-term performance of in situ reactive barriers for ni-
    trate remediation. Ground Water 38: 689-695 (2000).
Robertson,  W.D.,  N.  Yeung, PW. van Driel, and  PS.
    Lombardo.  High-permeability layers for remediation
    of ground water; go wide, not deep. Ground Water 43:
    574-581  (2005).
Schipper,  L.A., A.B. Cooper, and  W.J.  Dyck.  Mitigating
    nonpoint-source nitrate pollution  by riparian-zone
    denitrification.  In  Nitrate contamination:  Exposure,
    consequence,  and control. NATO ASI Series G: Eco-
    logical Sciences 30,1. Bogardi and R.D. Kuzelka (Eds.),
    Springer-Verlag, Berlin,  pp. 401-413 (1991).
Schnabel,  R.R., L.F Cornish, W.L. Stout, and J.A. Shaffer.
    Denitrification in grassed and a wooded,  Valley and
    Ridge,  riparian ecotone. Journal of Environmental
    Quality 25: 1230-1235 (1996).
Schuman,  G.E. and L.F. Elliott.  Cropping an abandoned
    feedlot to prevent deep percolation of nitrate-nitrogen.
    So/7 Science 126: 237-243 (1978).
Schurmann,  A., M.H. Schroth, M. Saurer, S.M. Bernasconi,
    and J.  Zeyer. Nitrate-consuming processes in a petro-
    leum-contaminated aquifer quantified using push-pull
    tests combined with 15N isotope and acetylene-inhibi-
    tion methods.  Journal of Contaminant Hydrology 66:
    59-77  (2003).
Sigman, D.M., K.L. Casciotti, M. Andreani, C. Barford, M.
    Galanter, and  J.K. Bohlke.  A bacterial method  for
    the  nitrogen isotopic analysis of nitrate in seawater
    and freshwater. Analytical Chemistry 73:  4145-4153
    (2001).
                                                     98

-------
Skaggs, R.W. and J.W. Gilliam. Effects of drainage system
    design and operation on nitrate transport. Transactions
    of American Society of Agricultural Engineers 24: 929-
    934(1981).
Smith, R.L., S.R  Garabedian, and M.H. Brooks.  Com-
    parison  of denitrification activity  measurements in
    groundwater using  cores and natural gradient  tracer
    tests.  Environmental Science  and Technology 30:
    3448-3456(1996).
Smith, R.L., D.N. Miller, M.H. Brooks, M.A. Widdowson,
    and M.W. Killingstad.  In situ stimulation of ground-
    water denitrification with formate to remediate nitrate
    contamination. Environmental Science and Technology
    35: 196-203(2001).
S0rensen, J. and L. Thorling. Stimulation by lepidocroc-
    ite (y-FeOOH) of  Fe(ll)-dependent nitrite reduction.
    Geochimica  et  Cosmochimica Acta 55: 1289-1294
    (1991).
Spalding, R.F. and M.E. Exner.  Occurrence of  nitrate in
    groundwater-A review. Journal of Environmental Quality
    22:392-402 (1993).
Squillace, P.J., J.C. Scott, M.J. Moran, B.T. Nolan, and D.W.
    Kolpin.  VOCs, pesticides, nitrate, and their mixtures
    in groundwater used for drinking water in the United
    States.  Environmental Science and Technology 36:
    1923-1930(2002).
Strathouse, S.M., G. Sposito, P.J. Sullivan, and L.J. Lund.
    Geological nitrogen: A potential geochemical hazard
    in the San Joaquin Valley, California. Journal of Envi-
    ronmental Quality 9: 54-60 (1980).
Strock, J.S., P.M. Porter, and M.R Russelle. Cover cropping
    to reduce nitrate loss through subsurface drainage in
    the Northern U.S. corn belt. Journal of Environmental
    Quality 33: 1010-1016 (2004).
Su, C. and R.W. Puls. Nitrate reduction by zerovalent iron:
    Effects  of formate,  oxalate,  citrate, chloride, sulfate,
    borate,  and phosphate.  Environmental Science  and
    Technology 38: 2715-2720 (2004).
Tesoriero, A.J., H. Liebscher, and S.E. Cox. Mechanism and
    rate of denitrification in an agricultural watershed: Elec-
    tron and mass balance along groundwater flow  paths.
    Water Resources Research 36:1545-1559 (2000).
Tezuka, S.,  R. Chitrakar, A. Sonoda, K. Ooi, and T. Tomida.
    Studies on selective adsorbents for oxo-anions. Nitrate
    ion-exchange properties of layered double hydroxides
    with different  metal atoms. Green Chemistry 6: 104-
    109 (2004).
Tiedje, J.M. Denitrifiers.  In Methods of soil analysis, part
    2, 3rd ed., SSSA Book Series 5,  R.W. Weaver et al.
    (Eds.), Soil Science Society of America, Madison, Wl,
    pp. 245-268(1994).
Trolard, F,  J.M.R. Genin, M. Abdelmoula, G. Bourrie, B.
    Humbert, and A. Herbillon.  Identification of a  green
    rust mineral in a reductomorphic soil by Mossbauerand
    Raman spectroscopies. Geochimica et Cosmochimica
    Acta 61: 1107-1111 (1997).
Trudell,  M.R., R.W. Gillham, and J.A. Cherry.  An in situ
    study of the occurrence and rate of denitrification in a
    shallow unconfined sand aquifer. Journal of Hydrology
    83:251-268(1986).
USEPA.   Workshop  on Monitoring Oxidation-Reduction
    Processes for Ground-water Restoration, EPA/600/R-
    02/002, Office of  Research and Development, Cincin-
    nati, OH (2002).
USEPA. 2006 Edition of the Drinking Water Standards and
    Health Advisories. EPA 822-R-06-013, Office of Water,
    Washington DC, (2006).
van Driel, P.W., W.D. Robertson, and L.C. Merkley. Upflow
    reactors for riparian zone denitrification.  Journal of
    Environmental Quality 35: 412-420 (2006).
Vidon, P.  and A.R. Hill.  Denitrification and patterns of
    electron donors and acceptors in eight riparian zones
    with contrasting  hydrogeology. Biogeochemistry 71:
    259-283 (2004).
Vogel, J.C., A.S. Talma, and T.H.E. Heaton. Gaseous ni-
    trogen  as evidence for denitrification in groundwater.
    Journal of Hydrology 50: 191 -200 (1981).
Walvoord, M.A., FM. Phillips, D.A. Stonestrom, R.D. Evans,
    PC. Hartsough,  B.D.  Newman, and R.G. Striegl.  A
    reservoir of nitrate beneath desert soils. Science 302:
    1021-1024(2003).
Widory,  D., E. Petelet-Giraud, P. Negrel, and B. Ladouche.
    Tracking the sources  of nitrate in groundwater using
    coupled nitrogen  and boron  isotopes:  A  synthesis.
    Environmental Science and Technology 39:  539-548
    (2005).
Williamson, A.K., M.D. Munn, S.J. Ryker, R.J.  Wagner,
    J.C. Ebbert,  and A.M. Vanderpool.  Water quality in
    the  Central Columbia  Plateau, Washington and Idaho,
    1992-1995. U.S. Geological Survey  Circular  1144,
    35pp. (1998).
Yang, G.C.C. and H.L. Lee.  Chemical reduction of nitrate by
    nanosized iron: Kinetics and pathways. Water Research
    39: 884-894 (2005).
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                                              Perch lorate
                        Patrick V. Brady, Richard T. Wilkin, Robert G. Ford
Occurrence and Distribution

Ammonium perchlorate  (NH4CIO4) has been used with
powdered aluminum and various chemical binders as rocket
fuel and in munitions since the mid-1940s. It powers the
space shuttle and the U.S. nuclear missile arsenal.  There
are no ready substitutes. Potassium perchlorate is used in
airbags in cars, in the production of leather, fabrics, color-
ing agents, fireworks (Wilkin et al., 2007),  and elsewhere.
Perchlorate in the environment is largely associated with the
periodic replacement of perchlorate in  rockets (ammonium
perchlorate has a long, but not infinite, shelf-life) and its
subsequent disposal.  Naturally-occurring perchlorate from
nitrate deposits in Chile has occasionally been released to
the environment during mining to recover nitrate for use as
fertilizer (e.g., Urbansky et al., 2001). Perchlorate has also
been detected in sylvite from New Mexico, as well as in lang-
benite - a potassium sulfate mineral. The most common co-
contaminants found at perchlorate-contaminated sites are
nitrate and sulfate (ITRC, 2005). Atmospheric sources of
perchlorate have also been implicated as a potential source
of perchlorate in arid and semi-arid regions (Rajagopalan et
al., 2006). Perchlorate is a powerful oxidant when heated,
but the  oxidation reaction is  kinetically sluggish at low
temperatures. The combination of low temperature stability
and high temperature reactivity is what makes perchlorate
so attractive for explosives and rocket fuel.

Plume Characteristics

Because perchlorate is a large anion  with a relatively low
diffuse  charge it  is non-complexing,  forms no insoluble
minerals, and sorbs poorly to most  solids. The solubility of
perchlorate salts  is typically in the range of 10 - 2000 g  L1.
Although perchlorate sorption  to soil materials is typically
assumed to be negligible, there is some evidence that
perchlorate might sorb to soil organics  (e.g., Urbansky and
Brown, 2003). Partitioning coefficients (i.e., Kd) describing
sorption to geologic materials are  usually found to equal
zero. Few ion exchange resins or filter media effectively
remove  perchlorate from drinking water  or wastewater.
Similarly, perchlorate sorbs poorly to most geologic materi-
als, suggesting that perchlorate plumes would likely move
at roughly the same speed as ground water in the absence
of biodegradation.  Perchlorate-rich brines released from
industrial applications can be sufficiently dense as to 'sink'
in the subsurface analogous to dense  non-aqueous phase
liquids (Flowers and Hunt, 2000; Motzer, 2001).
Remedial Technologies

A current review of available technologies for remediation
of  perchlorate contamination  in ground  water has  been
published by the Interstate Technology Regulatory Council
(ITRC, 2005), which is available via the internet  at http://
www.itrcweb.org/Documents/PERC-1.pdf. Available tech-
nologies include:  1) ex-situ, pump-and-treat approaches
that extract perchlorate from  recovered  ground  water or
degrade perchlorate to  innocuous compounds  through
biotic  or abiotic chemical reduction, and 2) in-situ treat-
ment  approaches that degrade perchlorate  within the
plume via stimulation of native microbial communities or
the emplacement  of media that achieve biotic and abiotic
chemical reduction of perchlorate, respectively. The suc-
cess of these treatment approaches will be influenced by
the level of knowledge of the ground-water flow regime as
well as the geochemical characteristics within  the plume.
For technologies based on perchlorate extraction, the pres-
ence of common anions in ground water may compete for
binding sites on ion exchange resins or other solid media
used  in the treatment process. For microbial perchlorate
reduction (ex-situor in-situ),  geochemical parameters such
as low pH and high total  dissolved solids, as well as the
presence of aerobic conditions or competing electron ac-
ceptors such as nitrate or sulfate may negatively impact
the rates and extent of perchlorate reduction that can be
achieved.  It is important to consider these site-specific
factors when selecting treatment technologies.

Regulatory Aspects

Introduction  of a  new analytical  method in 1997 by the
California Department of Health Services lowered the de-
tection limit for perchlorate from 400 ppb to 4 ppb.  Since
that time relatively wide occurrences of perchlorate have
been detected. As of 2001,  forty-four states have perchlo-
rate manufacturers or users (Motzer, 2001). There were
24 sites on  the National Priorities List  with perchlorate
contamination as  of September 23, 2004 (USEPA, 2004)
with various cleanup actions  having been initiated at 12
of  these sites  (http://www.epa.gov/swerffrr/documents/
perchlorate_site_summaries.htm). Perchlorate has  been
found in the water supplies of over 15 million people in
California, Nevada and Arizona and in surface or ground
water throughout the United States.

Perchlorate is used to treat Grave's Disease (hyperthyroid-
ism) and in sufficient quantity  competes with iodide in the
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thyroid gland.  The perchlorate ion (CIO4") has a charge
and radius close to iodide and substitutes for iodide in the
thyroid causing a decrease in hormonal output. Perchlorate
is apparently not metabolized in the body, nor is it rapidly
reduced to chloride in many soil settings.

There are at present no state or federal drinking water
standards for perchlorate, but several  are being consid-
ered.  With authority under the Safe Drinking Water Act
(SDWA), in March 1998, the  U.S. Environmental Protec-
tion Agency's Office of Water formally  added perchlorate
to  the drinking water contaminant candidate list  (CCL)
(Urbansky and Schock 1999).  A  National Primary Drink-
ing Water Regulation (NPDWR) has not yet been promul-
gated for perchlorate, but monitoring will  continue under
the Unregulated Contaminants Monitoring Rule  (UCMR).
Several  states have set  action or advisory levels for per-
chlorate: Nevada -18 ppb; Arizona -14 ppb; California and
Texas - 4 ppb; Maryland and Massachusetts - 1 ppb. No
state governments have set enforceable standards. Cur-
rent efforts are focused  on identifying  the specific  health
risks posed by perchlorate  intake  (National  Research
Council,  2005; http://www.nap.edu/catalog/11202.html).
Perchlorate compounds do not have  reportable quanti-
ties under the Comprehensive Environmental Response,
Compensation, and Liability Act (CERCLA), as amended,
but the  USEPA Office of Solid Waste and  Emergency
Response has issued recent guidance with a Drinking
Water Equivalent  Level (DWEL) for actions administered
under the CERCLA of 24.5 ug L~1 based on  an adopted
reference dose (RfD) for perchlorate of 0.0007 mg/kg-day
(http://www.epa.gov/fedfac/pdf/perch lorate_guidance.pdf).

Geochemistry and Attenuation Processes

Attenuation  of  perchlorate might be  achieved through
reduction ultimately to chloride.  Specialized microorgan-
isms able to reduce perchlorate to  chloride have  been
identified (Coates and Achenbach, 2004; Achenbach et
al., 2001; Coates  et al.,  1999) that are widespread in the
environment.  All of  the identified perchlorate  reducers
are facultatively anaerobic or microaerophilic (Coates et
al., 1999). Perchlorate-reducing bacteria use as electron
donors:  hydrogen, organic acids  and  alcohols,  aromatic
hydrocarbons, reduced humic substances, hexoses, ferrous
minerals and hydrogen  sulfide (Chaudhuri et al., 2002).
In  situ perchlorate biodegradation methods typically use
electron donors such as vegetable oils, organic  acids,
alcohols, or sugars. Anaerobic conditions are required for
biodegradation of perchlorate to proceed. Metabolism of
electron donors leads to the consumption of free oxygen
and nitrate, at which point perchlorate reduction proceeds.
Molybdate is apparently important for perchlorate reduction;
nitrate occasionally is seen to inhibit perchlorate reduction
(Chaudhuri et al.,  2002).

Dozens  of strains of perchlorate-reducing bacteria have
been identified in the environment, their ubiquity, substrate
specificity,  and wide  distribution suggesting both  a
natural source of perchlorate and the potential for natural
attenuation of perchlorate. Perchlorate reduction to chloride
proceeds in the general sequence of perchlorate to chlorate
to chlorite to hypochlorite to chloride:

CIO4- - CIO3- - CIO2- - CIO- - Cl- + O2     (Eq. 9.1)

The first step of perchlorate reduction to chlorite is thought
to be the rate-limiting step in the cascade.

The Chilean Perchlorate Deposits

Two lines of evidence suggest that natural perchlorate fluxes
into soils might be appreciable and that  natural attenuation
of very small amounts of perchlorate is perhaps common.
The specificity and relative ubiquity of  perchlorate reduc-
ing microorganisms imply the existence of an undetected,
"shadow" flux of perchlorate (Urbansky and Schock, 1999).
As noted by Urbansky and Schock (1999),

   "If it can be shown  that perchlorate is produced
   naturally in the environment and yet levels are very
   low,  we must conclude that natural attenuation is
   responsible for the dichotomy."

Presumably, the development of more  powerful analytical
methods will point to the existence or non-existence of
such a flux.

The occurrence of natural perchlorate in the nitrate depos-
its of the Atacama  Desert in  Chile points  to the existence
of a  natural process  of perchlorate formation - though it
says little more.   Perchlorate minerals occur along with
other oxidized minerals rarely observed elsewhere includ-
ing nitrates, iodates,  chromates and dichromates.  There
are competing explanations for the origin(s) of the nitrate
salts - which have  received the most attention - including
atmospheric formation (through electrochemical or pho-
tochemical means) followed by evaporative concentration
(Bohlke et al., 1997; Dasgupta et al., 2005)  or fixation by
microorganisms (Ericksen, 1983). Ericksen (1983) sug-
gested a photochemical oxidation of atmospheric Cl and O
in the atmosphere or at ground level to form the perchlorate
deposits, citing as evidence arguments  that a similar reac-
tion might take place in the stratosphere.  The iodates in
Chile might have come from sea spray (much of the iodine
in the sea is iodate). The formation  path for chromate is
less  clear as there are few sources  of Cr near the  area.
The dichromate mineral  lopezite - K2Cr2O7  - appears to
be an alteration product  of the chromate minerals (Erick-
sen,  1983). Note that the half cell reaction for reduction of
perchlorate to chloride is:

CIO4- + 8e- + 8H+ -> Cl- + 4H2O   E0 =  1.29V   (Eq. 9.2)

which means that in acidic solutions perchlorate  is in
theory a stronger oxidant than oxygen, but not as strong
as dichromate (Urbansky and Schock, 1999).  In  other
words, a dichromate mineral once formed might be able to
oxidatively form perchlorate from chloride. In the absence
of a clearer understanding of how perchlorate came to form
in the Atacama Desert,  two poorly constrained potential
fluxes present themselves:
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    1.  An undefined atmospheric source of perchlo-
       rate - perhaps oxidation of chloride to chlorate
       by OH radicals, followed by further photochemi-
       cal oxidation by H2O2, ozone or HO2  radical
       (e.g., Dasgupta et al., 2005),
    2.  An undefined soil source of perchlorate linked
       to oxidation of chloride by  possibly transient
       dichromate sources or other unidentified soil
       processes.
Obviously the above is largely speculation in the absence of
a better understanding of mechanisms. Presumably, better
analytical methods will allow more  precise measurement
of actual  perchlorate backgrounds and fluxes.

Site Characterization

Site characterization, conceptual evaluation of long-term
stability and capacity, and tiered  analysis for perchlorate
is likely be similar to that done for other contaminants that
naturally  attenuate through reduction by ambient microor-
ganisms  - namely chlorinated solvents.  Anaerobic condi-
tions (nitrate reducing or lower) are critical as are sufficient
quantities of potential electron donors.   In the absence of
further research demonstrating otherwise, it is reasonable
to assume the presence of perchlorate-reducing bacteria
(Wu et al., 2001). A recent publication outlines an approach
used to assess the limitations for natural biodegradation of
perchlorate at a contaminated site (Kesterson et al., 2005).
These authors collected data on the presence of perchlo-
rate-reducing bacteria, availability of electron donors, and
the concentrations of competing electron acceptors. While
the presence of perchlorate-reducing bacteria in soil and
water was determined throughout the site, perchlorate
persisted in the subsurface due either to the lack of avail-
able electron donors and/or the  presence of competing
electron  acceptors.  This work illustrates the importance
of collecting site specific data to  explicitly demonstrate
that natural attenuation of perchlorate by biodegradation is
both active and of sufficient magnitude to achieve required
mass/concentration reductions in ground water. A general
summary is presented in Table 9.1 of the attenuation pro-
cess anticipated to dominate perchlorate transport in ground
water along with a synopsis of the types of information to
be collected to support site characterization.
                               Determining Cl Ground-water Spec/at/on

                               Determination of the potential extent of perchlorate deg-
                               radation within a plume can be initially assessed through
                               observation of the mass distribution of chlorine-bearing
                               inorganic ions according to the sequential reaction scheme
                               shown in Equation 9.1.   ITRC (2005) provides a review
                               of several published methods for determining perchlorate
                               or the distribution of Cl-bearing inorganic anions in water
                               samples, including those similar to published EPA methods
                               (Table 9.2). Unlike methods based on analyte detection by
                               conductivity, methods employing mass spectrometry are
                               capable of confirming the chemical composition of unknown
                               ions by their molecular weight (mass-to-charge ratio). This
                               may be important for samples with complex matrices con-
                               taining ions that may co-elute with perchlorate during ion
                               chromatography separation. Preservation studies suggest
                               that filtration and refrigeration (4C) of samples is sufficient
                               for  laboratory  analysis within reasonable holding times
                               (Stetson et al., 2006; Wilkin et al., 2007).  Comparison of
                               results from these studies suggests that filtration with a pore
                               size less than 0.45 u,m may be warranted to  eliminate the
                               potential  for microbial degradation  of perchlorate during
                               storage.  Since perchlorate  is a common component of
                               many detergents  used in field decontamination procedures
                               (such as Alconox, Alcotabs, Liqui-Nox, and Neutrad), it is
                               recommended that equipment rinseate blanks be collected
                               during each sampling event when sampling  equipment is
                               reused (ITRC, 2005). Thorne (2004) developed a field
                               colorimetric method that showed good agreement with EPA
                               Method 314.0 over the range of 1-225 ug L1 (slope = 1.11,
                               R2 = 0.913) for well water and bioreactor effluent samples.
                               The method is based on pre-concentration of perchlorate
                               onto a solid-phase extraction cartridge that has been condi-
                               tioned with a perchlorate-specific ion-pair reagent, followed
                               by elution into an ion-pairing dye that is further treated for
                               absorbance measurement at 640 nm using a standard
                               portable spectrophotometer.  No other published literature
                               was available to further document the field performance of
                               this method, but the general simplicity of the method sug-
                               gests that it may be suitable as a field screening tool when
                               there is an immediate need for analytical data to guide the
                               placement of additional  ground-water sampling locations
                               during initial site characterization efforts.
Table 9.1  Natural attenuation and mobilization pathways for perchlorate.
     Attenuation
      Processes
      Mobilization
       Processes
            Characterization Approach
 Biotic (microbial) or
 abiotic transformation
 of perchlorate to other
 chlorine species
Cessation of microbial
reduction processes via
changes in electron donor
or nutrient supply within the
ground-water flow path
Determination of the mass distribution of perchlorate and
other Cl-bearing inorganic ions in ground water.  Determi-
nation of water chemistry, microbial populations, and/or
sediment mineralogy linked to perchlorate transformation
along with spatial and temporal variability of required chemi-
cal/microbial components relative to perchlorate transport
pat h way (s).
                                                    103

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Table 9.2   Published USEPA methods for determination
           of perchlorate and other Cl-bearing inorganic
           Ions in aqueous samples.
Method Name and Number
Determination of Perchlorate in
Drinking Water by Ion Chroma-
tography; 374.0
Determination of Perchlorate
using Ion Chromatography with
Chemical Suppression Conduc-
tivity Detection; 9058, Revision 0
Perchlorate in Water, Soils and
Solid Wastes Using High Per-
formance Liquid Chromatogra-
phy/Electrospray lonization/Mass
Spectrometry; 6850, Revision 0
Perchlorate in Water, Soils and
Solid Wastes Using Ion Chroma-
tography/Electrospray lonization/
Mass Spectrometry
(IC/ESI/MS OR IC/ESI/MS/MS);
6860, Revision 0
Determination of Perchlorate
in Drinking Water Using Inline
Column Concentration/Matrix
Elimination Ion Chromatography
with Suppressed Conductivity
Detection; 374.7, Revision 1.0
Determination of Perchlorate in
Drinking Water by Liquid Chro-
matography Electrospray loniza-
tion Mass Spectrometry, 337.0,
Rev. 1.0
Determination of Perchlorate
in Drinking Water by Ion Chro-
matography with Suppressed
Conductivity and Electrospray
lonization Mass Spectrometry;
332.0
Determination of Inorganic
Anions by Ion Chromatography;
9056A, Revision 1
Source
November 1 999
http://www.epa.
gov/ogwdw/meth-
ods/pdfs/met314.
pdf
November 2000
http://www.epa.
gov/epaoswer/
hazwaste/test/
pdfs/9058.pdf
January 2007
http://www.epa.
gov/epaoswer/
hazwaste/test/
pdfs/6850.pdf
January 2007
http://www.epa.
gov/epaoswer/
hazwaste/test/
pdfs/6860.pdf
EPA815-R-05-
009, May 2005
http://www.epa.
gov/ogwdw/meth-
ods/pdfs/method
314_1.pdf
EPA815-R-05-
007, January 2005
http://www.epa.
gov/ogwdw/meth-
ods/pdfs/met331
O.pdf
EPA/600/R-
05/049, March
2005
http://www.epa.
gov/nerlcwww/
m_332_0.pdf
November 2000
http://www.epa.
gov/epaoswer/
hazwaste/test/
pdfs/9056a.pdf
Stable Isotope Techniques
Stable isotopes of chlorine and oxygen within the perchlo-
rate anion can be used to aid in differentiating natural and
synthetic sources of contaminant as well as the degree to
which perchlorate has been reduced by biotic processes.
Specifically, determination of the distribution of the stable
isotopes of oxygen (180, 170, 16O) can be used to determine
the value of A17O (or 17O anomaly), defined by the follow-
ing equations:
8-0 = (17CI/i6CI)sample/07CI/i6CI)reference- 1
A17O = [(1+
                       818O)0525]
                            0525 -
                                                                                                    (Eq. 9.3)

                                                                                                    (Eq. 9.4)

                                                                                                    (Eq. 9.5)
                                                       The value of A17O can be used to differentiate between natu-
                                                       ral sources of perchlorate (e.g., Atacama salt deposits and
                                                       fertilizer products) versus synthetic perchlorate produced by
                                                       electrolytic oxidation (Bao and Gu, 2004; Bohlke et al., 2005;
                                                       Motzer et al., 2006). As discussed by Bao and Gu (2004),
                                                       chemical reactions and biodegradation of perchlorate will
                                                       not change the value of A17O of the original perchlorate,
                                                       which can only occur within a ground-water plume through
                                                       addition of another source of perchlorate with a different
                                                       A17O signature.  In  addition, the values of 837CI  (defined
                                                       as rCI/^CI)   le/(37CI/35CI)reference- 1] relative to Standard
                                                       Mean  Ocean Chloride; Long et al., 1993) from Atacama
                                                       nitrate ore and from Chilean nitrate fertilizer products are
                                                       the lowest  reported for any common substance on Earth
                                                       (Bohlke et al., 2005).  These measurements have been
                                                       used to determine the sources of perchlorate in ground-
                                                       water plumes at several sites of contamination (Bohlke et
                                                       al., 2005).

                                                       Changes in the fractionation of the chlorine stable isotope
                                                       composition of perchlorate are significant during microbial
                                                       reduction (Coleman et al., 2003; Sturchio  et al., 2003),
                                                       consistent with preferential reduction of perchlorate con-
                                                       taining the lighter chlorine isotope (35CI). Coleman et al.
                                                       (2003) observed chlorine isotope fractionation on the order
                                                       of -15%0 during perchlorate reduction by a microorganism
                                                       isolated from a swine waste lagoon (Dechlorosoma suillum
                                                       strain PS). Sturchio et al. (2003) also assessed chlorine
                                                       stable  isotope fractionation during  perchlorate reduction
                                                       using Dechlorosoma suillum JPLRND isolated from ground
                                                       water sampled in southern California.  In addition, these
                                                       authors used EPA Methods 300.0 and 314.0 to monitor the
                                                       speciation of perchlorate,  chlorate,  chlorite, and chloride
                                                       throughout incubation  observing  only  minor  intermittent
                                                       concentrations of chlorate (<1.3 % of perchlorate concentra-
                                                       tion). The overall reaction was dominated by stoichiometric
                                                       conversion of  perchlorate to chloride.  Based  on these
                                                       experiments, Sturchio et al. (2003) estimated that chlorine
                                                       stable  isotope fractionation during  microbial  perchlorate
                                                       reduction may be detected at levels of biodegradation <2%,
                                                       which  may provide a more sensitive limit of detection than
                                                       that obtained through direct measurements of changes in
                                                       perchlorate and chloride concentrations within the ground-
                                                       water plume.
                                                     104

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Stable isotope measurements for this contaminant have
been  made possible through the development of a highly
selective bifunctional anion-exchange resin and sample pro-
cessing method to extract and concentrate perchloratefrom
environmental samples (Gu et al., 2000; Gu et al., 2001). As
reported by Bohlke et al. (2005), the collection of perchlorate
onto the anion-exchange resin can be accomplished in the
field with further processing conducted in the laboratory.
However, the acquisition of a sufficient mass of perchlorate
to support isotopic analyses for ground-water samples with
1-10 ug L1 perchlorate  may require significant volumes of
water (Duncan et al., 2005). Determination of the various
isotopes of chlorine and oxygen is dependent on the conver-
sion of perchlorate to various gaseous products that carry
the specific isotope of interest. For example, Bohlke et al.
(2005) describe the following procedures: 1) determination
of 818O following decomposition of perchlorate to carbon
monoxide, 2) determination of A17O following decomposition
of perchlorate to  produce oxygen, and 3) determination of
837CI  following decomposition of perchlorate to produce
chloride that was subsequently reacted to produce methyl
chloride (CH3CI; also known as chloromethane).  As these
authors point out, there are currently no perchlorate isotopic
reference  materials available for the purpose of assessing
performance for a given analytical facility.

Identifying Microbial Populations

Recent work has demonstrated the ubiquity of perchlorate-
reducing bacteria at sites with perchlorate contamination in
ground water (Waller et al., 2004). These microorganisms
reduced perchlorate in  microcosm studies with or without
supplement of an electron donor (e.g., acetate, molasses),
although degradation rates were greater for supplemented
microcosms.  In  order  to assess the degree of intrinsic
bioreduction of perchlorate at a given site,  information is
needed to identify the conditions of the  microbial process
controlling perchlorate reduction throughout the perchlorate
plume.  Bender et al. (2004) have developed two metabolic
primer sets designed to target the chlorite dismutase (eld)
gene, which can be used to identify perchlorate-reducing
bacteria in environmental samples by a qualitative poly-
merase chain  reaction (PCR) approach.  False positive
identification of perchlorate-reducing bacteria may occur,
since the approach targets only a single gene that may be
present even though the microorganism  lacks other genes
necessary to carry out perchlorate metabolism (e.g., genes
required for perchlorate reductase). However, subsequent
analysis of c/d-positive samples using perchlorate reductase
probes could be carried out in order to eliminate false posi-
tive identifications (Bender et al., 2005). This approach may
substitute for or supplement for enumeration of perchlorate-
reducing bacteria using the Most Probable Number (MPN)
procedure (Wu et al., 2001; Kesterson et al., 2005).

Long-term Capacity

Given reducing conditions and labile electron donor masses
that substantially exceed the perchlorate source, a critical
unknown for determining the long-term capacity for attenu-
ation  is likely to be the  perchlorate reduction rate relative
to the flux of  perchlorate and required electron donors
and essential nutrients to sustain the metabolic process.
Reduction rates for perchlorate remain an area of ongo-
ing research. There is no universally accepted means for
estimating perchlorate reduction rates,  although the use
of stable isotopes of chlorine and oxygen may provide a
means to estimate in-situ rates of reduction along a plume
flow path. Since there are potentially large uncertainties
associated with extrapolating degradation rates determined
from  laboratory microcosm  studies to the field setting,
multiple lines of evidence based on field observations will
likely  provide the  most reliable means  for assessing the
long-term capacity of the aquifer.  In general, a default ap-
proach might then be to: 1)  establish that conditions are
reducing; 2) demonstrate that there are more than sufficient
electron donors available for perchlorate reduction; 3) fit
plume contours over time to estimate a biodegradation rate
(possibly supported by microcosm studies demonstrating
that perchlorate reduction is possible under the conditions
prevailing in the field); and 4) use the derived model to as-
sess long-term  capacity for sustaining biological reduction
of perchlorate.

Tiered Analysis

Determination of the viability of perchlorate remediation in
ground water via monitored natural attenuation will depend
upon  proper assessment of contaminant loading to the
aquifer and prevailing geochemistry and mineralogy within
the contaminant plume and the down gradient zone prior to
the point(s) of compliance. It is essential to understand the
cause-and-effect relationship between loss of perchlorate in
ground water and the mechanisms responsible for its loss.
The following tiered analysis structure for site characteriza-
tion provides an approach to evaluate candidate sites and
define the potential limitations of MNA as part of a remedy
for ground-water cleanup.

Tier  I.  Site characterization under Tier I   will involve
demonstration that the plume is static  or shrinking, has
not reached compliance boundaries, and does not impact
existing water supplies.  Once this is established through
ground-water characterization, evidence is collected  to
demonstrate perchlorate reduction within the plume. Rapid
movement of contaminants along preferred flow paths in
the unsaturated and saturated zones can be  induced by
hydrologic events  such as heavy rains (e.g., McCarthy et
al., 1998; Camobreco et al., 1996). It will be important to
determine that such hydrogeologic features do  not result in
contaminants bypassing zones where natural  attenuation
is occurring. In addition, for sites in which the suspected
source term may have consisted of a perchlorate brine, it is
recommended that the hydrogeologic characterization of the
site establish whether perched 'DNAPL-like' plumes exist
that could influence both the timeframe for remediation and
the required attenuation capacity within the aquifer. If natural
attenuation processes are active throughout the  plume, then
an observed decrease in perchlorate concentration in space
and time that  is attributable to a mass-removal process
is anticipated.   Conditions that would  support this initial
screening would include evidence of reducing conditions
                                                    105

-------
(i.e., low EH), abundant electron donors (e.g., elevated total
organic matter or the presence of degradable organic co-
contaminants),  and the apparent availability of essential
nutrients such as molybdate. Identification of perchlorate
removal along ground-water flow paths provides justification
for proceeding to Tier II characterization efforts.

Tier II.  Under Tier II, the apparent rate and mechanism(s) of
attenuation are determined. Estimates of a site attenuation
rate(s) can be determined via several approaches: 1) as-
sessment of perchlorate disappearance across well tran-
sects along the ground-water flow path(s), 2) assessment
of in-situ rates of perchlorate reduction through analysis of
changes in the chlorine stabile isotope fractionation along
the ground-water flow path, and 3) ex-situ assessment of
perchlorate reduction rate(s) through the use of microcosm
studies. In addition, time-series data may be collected at
one or more monitoring points within the plume. This infor-
mation will allow assessment of the relative timescale(s) for
perchlorate reduction relative to ground-water seepage ve-
locities and determination of whether remediation objectives
can be met within the required regulatory time frame.  In
addition, the mechanism(s) for attenuation need to be iden-
tified under this stage of site characterization.  This effort
may require identification of the microbial agents involved
in perchlorate reduction and their distribution throughout the
plume  (see Site Characterization section in this chapter). If
a link is established between the apparent disappearance of
perchlorate and observed microbial processes active within
the plume, then this provides justification for proceeding to
Tier III characterization efforts.

Tier III.  Once the perchlorate reduction mechanism(s)
have been identified for the site, the subsequent charac-
terization effort under Tier III will involve determination of
the capacity of the aquifer to sustain continued biological
reduction. The impact of potential hydrologic changes, such
as a shift in flow direction caused by the onset of pumping
at nearby sites, needs to be determined.  Since the active
microbial community will reside within the boundaries of the
perchlorate plume, potential alterations in the predominant
flow path  may deliver perchlorate to regions of the aqui-
fer in which there is a reduced (or inactive) population of
perchlorate-reducing bacteria to sustain reduction. A well-
constrained ground-water flow model of the site relative to
observed  spatial distributions of  perchlorate and density
perchlorate-reducing bacteria will assist in assessing the
potential impact of such changes. For sites in which  or-
ganic co-contaminants are identified as the predominant
electron donor for perchlorate reduction, it is recommended
that an evaluation of the long-term supply of the electron
donor be assessed.  If engineered remedies for the treat-
ment  of co-contaminants are active  or planned (e.g.,
source zone  treatment or removal  of  co-contaminants),
then it is expected that  the impact these remedies have
on the potential supply of electron donor will influence the
sustainability  of perchlorate  reduction.  This may require
evaluation of whether the total organic carbon content and
characteristics of treated  ground water are sufficient to
maintain the activity of perchlorate reducing bactera. This
analysis could be conducted through laboratory microcosm
studies. For sites in which uncontaminated portions of the
aquifer are aerobic (i.e., oxygen is present) or have naturally
elevated concentrations of competing electron acceptors
such as nitrate, it is anticipated that evolution of the plume
geochemistry may evolve to a condition that will not sustain
microbial perchlorate reduction. It is recommended that the
potential  impact of  anticipated changes in ground-water
chemistry on the rate or sustainability of perchlorate reduc-
tion be assessed  through  microcosm studies.  Ultimately,
the ability to forecast the potential impacts of changes in
aquifer chemistry or ground-water flow can  be improved
through the development of a reaction-transport model
that includes site-specific parameterization of the microbial
reduction  process and site hydrogeology.  If site-specific
tests demonstrate sufficient  capacity within the aquifer to
sustain perchlorate  reduction, then the site characteriza-
tion effort can progress to Tier IV. For cases where aquifer
capacity is insufficient for plume reduction to required levels,
then a determination of the benefits of contaminant source
reduction or removal may be necessary.

Tier IV.  Finally, under Tier IV a monitoring plan is estab-
lished along with  contingency plans in the event of MNA
failure.  It is  recommended  that the monitoring plan  be
designed to establish both continued plume stability and to
identify changes in ground-water chemistry that may lead
to a decrease in the rates or capacity for perchlorate reduc-
tion. The specific chemical parameters to be monitored will
include those identified under Tier III that may halt or slow
down microbial perchlorate reduction within the plume. For
example, solution phase parameters  that could alter per-
chlorate reduction include the concentrations of dissolved
oxygen or competing electron acceptors such as nitrate.
Similarly, a decrease in the concentration of electron donors
such as natural organic matter or organic co-contaminants
could slow or halt  perchlorate biodegradation. Changes in
these parameters may occur prior to observed changes in
perchlorate concentrations and, thus, serve as  monitoring
triggers for potential MNA failure.  In this instance, a contin-
gency plan can be implemented that incorporates strategies
to arrest possible plume  expansion  beyond compliance
boundaries. Possible strategies to prevent plume expansion
include pump and treat operations, installation  of reactive
barriers to enhance uptake/degradation capacity perpen-
dicular to the direction of plume advance, or enhancement
of natural attenuation processes within the aquifer through
the injection of soluble reactive components (see Remedial
Technologies section in this  chapter).

References

Achenbach, L.A., R.A.  Bruce, U.  Michaelidou, and J.D.
    Coates.   Dechloromonas agitata N. N. gen. sp. nov.
    and Dechlorosoma suillum N. N. gen. sp. nov., two
    novel environmentally dominant (per)chlorate reducing
    bacteria and their phylogenetic position.  International
    Journal of Systematic and Evolutionary Microbiology
    51:527-533(2001).
                                                     106

-------
Bao, H. and B. Gu. Natural perchlorate has a unique oxygen
    isotope signature. Environmental Science and Technol-
    ogy 38: 5073-5077 (2004).
Bender, K.S., M.R. Rice, W.H.  Fugate, J.D. Coates, and
    L.A. Achenbach. Metabolic primers for detection of
    (per)chlorate-reducing  bacteria  in the  environment
    and phylogenetic analysis of eld gene sequences. Ap-
    plied and Environmental Microbiology 70: 5651 -5658
    (2004).
Bender, K.S., C.  Shang, R. Chakraborty,  S.M.  Belchik,
    J.D. Coates, and L.A. Achenbach. Identification, char-
    acterization,  and classification  of  genes encoding
    perchlorate  reductase. Journal of Bacteriology 187:
    5090-5096 (2005).
Bohlke, J.K., G.E. Ericksen, and K. Revesz. Stable isotope
    evidence for an atmospheric origin of  desert nitrate
    deposits in  northern Chile and  southern  California,
    USA. Chemical Geology-\ 36: 135-152 (1997).
Bohlke, J.K., N.C. Sturchio, B. Gu, J. Horita, G.M. Brown,
    W.A.  Jackson, J. Batista, and RB. Hatzinger.   Per-
    chlorate isotope forensics.   Analytical Chemistry 77:
    7838-7842 (2005).
Camobreco, V.J., B.K. Richards, T.S. Steenhuis, J.H. Peverly,
    and M.B. McBride. Movement of heavy metals through
    undisturbed and homogenized soil columns.  So/7 Sci-
    ence 161: 740-750 (1996).
Chaudhuri, S.K.,  S.M.  O'Connor, R.L. Gustavson, L.A.
    Achenbach, and J.D. Coates. Environmental factors
    that control microbial perchlorate reduction. Applied and
    Environmental Microbiology 68: 4425-4430 (2002).
Coates, J.D. and L.A. Achenbach.  Microbial perchlorate
    reduction: rocket-fuelled metabolism. Nature  Reviews
    Microbiology 2: 569-580 (2004).
Coates, J.D., U. Michaelidou, R.A.  Bruce, S.M.  O'Connor,
    J.N. Crespi, and L.A. Achenbach. The ubiquity and di-
    versity of dissimilatory (per)chlorate-reducing bacteria.
    Applied and Environmental Microbiology 65:5234-5241
    (1999).
Coleman,  M.L.,  M. Ader, S. Chaudhuri, and J.D. Coates.
    Microbial isotopic fractionation of perchlorate chlorine.
    Applied and Environmental Microbiology  69:  4997-
    5000 (2003).
Dasgupta, P.K., PK. Martinelango, W.A. Jackson, T.A. An-
    derson, K. Tian, R.W. Tock, and S. Rajagopalan. The
    origin  of naturally occurring perchlorate: The role of
    atmospheric processes.  Environmental Science and
    Technology 39: 1569-1575 (2005).
Duncan, P.B., R.D. Morrison, and E. Vavricka. Forensic
    identification of anthropogenic and naturally occurring
    sources of perchlorate. Environmental  Forensics 6:
    205-215(2005).
Ericksen,  G.E. The Chilean nitrate deposits.  American
    Scientist 71: 366-374 (1983).
Flowers, T.C. and J.R. Hunt. Long-term release of perchlorate
    as a potential source of groundwater contamination. In
    Perchlorate in the  Environment. E. Urbansky (Ed.),
    Kluwer/Plenum, New York, NY, pp. 177-187 (2000).
Gu, B., G.M. Brown, S.D. Alexandratos, R. Ober, J.A. Dale,
    and S. Plant. Efficient treatment of perchlorate (CIO4-)-
    contaminated groundwater  with bifunctional anion-
    exchange resins.  Environmental Science Research
    57: 165-176(2000).
Gu, B.H., G.M.  Brown,  L. Maya, M.J. Lance, and B.A.A.
    Moyer.  Regeneration of perchlorate (CIO4')-loaded
    anion-exchange resins by a novel tetrachloroferrate
    (FeCI4-) displacement technique. Environmental Sci-
    ence and Technology 35: 3363-3368 (2001).
ITRC (Interstate Technology  & Regulatory Council).
    Perchlorate: Overview of Issues, Status, and Remedial
    Options. PERCHLORATE-1. Washington DC: Interstate
    Technology & Regulatory Council, Perchlorate Team,
    152pp. (2005).
Kesterson, K.E., PS. Amy, and J.R. Batista. Limitations to
    natural bioremediation of perchlorate in a contaminated
    site. Bioremediation Journal 9: 129-139 (2005).
Long, A., C.J. Eastoe, R.S. Kaufmann, J.G. Martin, L. Wirt,
    and J.B. Finley. High-precision measurement of chlorine
    stable isotope ratios. Geochimica  et Cosmochimica
    Acta 57: 2907-2912 (1993).
McCarthy, J.F, K.R.Czerwinski, W.E.Sanford, RM.Jardine,
    and J.D. Marsh. Mobilization of transuranic  radionu-
    clides from disposal trenches by natural organic matter.
    Journal of Contaminant Hydrology 30: 49-77  (1998).
Motzer, W.E. Perchlorate: problems, detection, and  solu-
    tions. Environmental Forensics 2: 301 -311 (2001).
Motzer,  W.E., T.K.G. Mohr,  S.  McRaven, and P. Stanin.
    Stable and  other isotope techniques for perchlorate
    source identification.  Environmental Forensics 7: 89-
    100(2006).
National Research Council. Health Implications of  Perchlo-
    rate Ingestion, National Academies Press: Washington
    DC, 276 pp. (2005).
Rajagopalan, S.,T.A. Anderson, L. Fahlquist, K.A. Rainwater,
    M. Ridley, and W.A. Jackson. Widespread presence
    of naturally  occurring perchlorate  in  high plains of
    Texas and New Mexico.  Environmental Science and
    Technology 40: 3156-3162 (2006).
Stetson, S.J., R.B. Wanty, D.R. Helsel,  S.J. Kalkhoff, and
    D.L. Macalady. Stability of low levels of perchlorate in
    drinking  water and  natural  water samples. Analytica
    Chimica Acta 567: 108-113  (2006).
Sturchio, N.C., RB. Hatzinger, M.D. Arkins, C. Suh, and L.J.
    Heraty. Chlorine isotope fractionation during microbial
    reduction of perchlorate.  Environmental Science and
    Technology 37: 3859-3863 (2003).
Thorne, RG. Field Screening Method for Perchlorate in
    Water and Soil. ERDC/CRRELTR-04-8. Hanover,  N.H.:
    U.S. Cold Regions Research and Engineering Labora-
    tory, 36 pp.  (http://www.crrel.usace.army.mil/techpub/
    CRREL_Reports/reports/TR04-8.pdf) (2004).
                                                    107

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Urbansky, E.T., S.K. Brown, M.L. Magnuson, and C.A. Kelty.
    Perchlorate levels in samples of sodium nitrate fertilizer
    derived from Chilean caliche. Environmental Pollution
    112:299-302 (2001).
Urbansky, E.T. and M.R. Schock. Issues in managing the
    risks associated  with  perchlorate  in drinking  water.
    Journal of Environmental  Management 56:  79-95
    (1999).
Urbansky, E.T. and S.K. Brown. Perchlorate retention and
    mobility in soils. Journal Environmental Monitoring 5:
    455-462 (2003).
USEPA. National Perchlorate Detections as of September
    23, 2004.  Federal Facilities Restoration and Reuse
    Office,  Available at www.epa.gov/swerffrr/documents/
    perchlorate_map/nationalmap.htm.  (2004).
Waller, A.S.,  E.E. Cox, and E.A. Edwards.  Perchlorate-
    reducing microorganisms isolated from contaminated
    sites. Environmental Microbiology 6: 517-527 (2004).
Wilkin, R.T., D.D. Fine, and N.G. Burnett. Perchlorate be-
    havior in a municipal lake following  fireworks displays.
    Environmental Science and Technology 41:3966-3971
    (2007).
Wu, J., R.F. Unz, H. Zhang, and B.E. Logan. Persistence
    of perchlorate and the relative numbers of perchlorate
    and chlorate respiring  microorganisms in natural wa-
    ters, soils, and wastewater. Bioremediation Journal 5:
    119-130(2001).
                                                    108

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