United States
Environmental Protection
Agency
Monitored Natural
of Inorganic Contaminants in
Ground Water
Volume 1
Technical Basis for Assessment
Evolution of Inorganic Contaminant Plume
Timel
Mobile plume
shrinkage due to
degradation or
immobilization
onto aquifer solids
Immobilized
inorganic
contaminant still
present on
aquifer solids
Original Plume
Boundary
Mobile
Contaminant
Itttia
Immobile
Contaminant
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EPA/600/R-07/139
October 2007
Monitored Natural Attenuation of Inorganic
Contaminants in Ground Water
Volume 1 - Technical Basis for Assessment
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 the research 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 environmental problems today and building a science knowledge base
necessary to manage our ecological resources wisely, understand how pollutants affect our health, and
prevent or reduce environmental risks in the future.
The National Risk Management Research Laboratory is the Agency's center for investigation of technologi-
cal and management approaches for preventing and reducing risks from pollution that threatens human
health and the environment. The focus of the Laboratory's research program is on methods and their
cost-effectiveness for prevention and control of pollution to air, land, water, and subsurface resources;
protection of water quality in public water systems; remediation of contaminated sites, sediments and
ground water; prevention and control of indoor air pollution; and restoration of ecosystems. NRMRL
collaborates with both public and private sector partners to foster technologies that reduce the cost of
compliance and to anticipate emerging problems. NRMRLs research provides solutions to environmental
problems by: developing and promoting technologies that protect and improve the environment; advanc-
ing 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 commu-
nity 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 char-
acterization approaches useful for evaluating the effectiveness of the natural attenuation component of
ground-water remedial actions.
G. Schmelling, Director
Ground Water and Ecosystems/ftes/oration Division
National Risk Management Research Laboratory
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Contents
Notice ii
Foreword iii
Figures viii
Tables ix
Acknowledgments x
Executive Summary xi
Section I - Conceptual Background for Natural Attenuation
I A. Background and Purpose 1
IA.1 Document Organization 1
IA.2 Purpose of Document 1
IA.3 Applicable Regulatory Criteria 2
IA.4 Policy Framework for Use of MNA 2
IB. Relevant Distinctions in Site Characterization for MNA of Inorganic Contaminants 4
1C. Tiered Analysis Approach to Site Characterization 5
IC.1 Tier 1 6
IC.2 Tier II 7
IC.3 Tier III 7
IC.4 Tier IV 8
ID. Role of Modeling in the Tiered Analysis Approach 10
ID.1 Developing a Conceptual Model 10
ID.2 Types of Models 11
ID.2.1 Simple Calculations 11
ID.2.2 Mass Transport Models 12
ID.2.3 Speciation Models 12
ID.2.4 Reaction Models 12
ID.2.5 Reactive Transport Models 12
ID.3 Modeling and the Tiered Analysis Approach 13
ID.3.1 Tier I - Demonstration of Contaminant Removal from Ground Water 13
ID.3.2 Tier II - Determine Mechanism and Rate of Attenuation 13
ID.3.3 Tier III - Demonstrate Capacity and Stability of Removal Mechanism 13
ID.3.4 Tier IV - Long-Term Performance Monitoring 14
ID.4 Choosing Modeling Software 14
ID.4.1 Public Domain vs. Commercial Software 14
ID.4.2 Sources of Software 14
ID.4.3 Thermodynamic Data 16
ID.5 Accounting for Uncertainty 17
ID.6 Model Calibration and Verification 17
IE. Long-Term Performance Monitoring and Site Closure 18
IE.1 Duration and Monitoring Frequency 19
IE.2 Monitoring of Aquifer Solids 19
IE.3 Monitoring Types 19
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IE.4 Monitoring Locations 20
IE.5 Modification of the Performance Monitoring Plan 22
IE.6 Periodic Reassessment of Contaminant Removal Technologies 22
IF. References 23
Section II - Technical Basis for Natural Attenuation in Ground Water
IIA. Physical Transport Mechanisms 26
IIA.1 Basics of Ground-Water Flow and Solute Movement 26
IIA.2 Colloidal Transport of Inorganic Contaminants 27
IIA.2.1 Implications for Monitored Natural Attenuation 27
MB. Contaminant Sorption to Aquifer Solids 28
IIB.1 Adsorption 29
MB.1.1 Reactive Mineral Phases Involved in Adsorption 29
IIB.1.2 Surface Functional Groups on Aquifer Solids and the Impact on Surface
Charge 31
MB.1.3 Weak and Strong Adsorption Regimes 32
MB.2 Precipitation 33
MB.2.1 Precipitation from Solution 33
MB.2.2 Coprecipitation 35
MB.2.3 Surface Precipitation 35
MB.2.4 Mineral Transformation 36
MB.3 Implications for Natural Attenuation Assessment 36
IIC. Microbial Impacts on Inorganic Contaminant Attenuation 36
IIC.1 Characteristics of Aquifer Microbiology 37
IIC.2 Microbial Controls on Subsurface Redox State 37
IIC.3 Impacts on Contaminant Speciation and Attenuation 39
IIC.3.1 Contaminant Oxidation-reduction Reactions 39
IIC.3.2 Biosorption and Intracellular Bioaccumulation 39
IIC.3.3 Methylation and Demethylation 40
IIC.4 Implications for Natural Attenuation Assessment 40
I ID. References 40
Section III - Site Characterization to Support Evaluation of MNA
IIIA.Site Hydrogeology 43
I IIA.1 Characterization Objectives 43
IIIA.2 Geologic Characterization 44
IIIA.2.1 Saturated Porous Media 44
IIIA.2.2 Saturated Fractured Media 44
IIIA.3 Hydrologic Characterization 45
IIIA.4 Ground-Water/Surface-Water Interactions 47
IIIA.5 Hydrogeologic Data Interpretation 47
IIIA.5.1 Attenuation Rate Estimates 48
IIIA.5.2 Contaminant Flux 48
IIIA.5.3 Source Term Characteristics 50
NIB.Contaminant Quantification, Distribution and Speciation 50
IIIB.1 Aqueous Characterization Approaches 50
IIIB.1.1 Filtration 51
IIIB.2 Solid Phase Characterization Approaches 52
IIIB.2.1 Sampling and Fractionation 53
IIIB.2.2 Total Amount 53
IIIB.2.3 Structurally Defined Form 54
IIIB.2.4 Operationally Defined Form 55
IIIB.2.4.1 Sequential Extractions 55
VI
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IIIB.2.4.2 Sequential Extraction Considerations 58
IIIB.2.5 Attenuation Capacity 59
IIIB.3 Model Representations to Interpret Contaminant Sorption Observations 61
NIB.3.1 Distribution Coefficient/partition Coefficient, Kd 61
NIB.3.2 The Langmuir Model 61
IIIB.3.3 The Freundlich Isotherm 61
NIB.3.4 Mechanistic Models for Predicting Sorption - Surface Complexation 62
IIIB.3.5 Mineral Solubility 63
NIB.3.5.1 Coprecipitation Reactions 64
NIB.3.5.2 Thermodynamic Data 65
IllC.Characterization of System Redox and Underlying Microbial Processes 66
IIIC.1 Process Identification 66
IIIC.1.1 Redox Measurements 67
IIIC.2 Capacity 68
IIIC.3 Stability 69
IIIC.4 Microbial Community Characterization 70
IIIC.4.1 Standard and Emerging Techniques 70
IIIC.4.2 Molecular Characterization 70
IIIC.4.3 Sampling Considerations 71
IIIC.5 Implications for Natural Attenuation Assessment 71
MID. References 72
VII
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Figures
Figure 1.1 Conceptual distinction between organic versus inorganic contaminant plume
behavior where natural processes are active within the ground-water aquifer 5
Figure 1.2 Conceptual depiction of the data collection effort to demonstrate whether sorption
to aquifer solids attenuates contaminant transport in ground water 6
Figure 1.3 Example of a network design for performance monitoring, including target zones
for monitoring effectiveness with respect to specific remedial objectives 15
Figure 2.1 Conceptual view of attenuation as the interaction of the contaminant with aquifer
constituents to form a product resulting in attenuation 25
Figure 2.2 Cross-sectional view of differences in solute migration due to differences in hydraulic
conductivity with accompanying differences in ground-water velocity and the
spreading of the solute front caused by dispersion 26
Figure 2.3. Representation of an aquifer mineral surface with (a) an outer-sphere surface
complex; (b) an inner-sphere surface complex; (c) a multinuclear surface complex or
a surface precipitate; and (d) absorption, or solid state diffusion and substitution of
the sorbate in the mineral structure 28
Figure 2.4 Examples of contaminant-specific sorption processes that may lead to attenuation
of the ground-water plume 29
Figure 2.5 Diagrammatic sketch of the structure of 1:1 and 2:1 phyllosilicate minerals 30
Figure 2.6 Surface charge of some hydroxides from pH 2 to 10 measured in different
electrolyte solutions shown in parentheses; positive and negative surface charge
shown above and below the x-axis, respectively. 31
Figure 3.1 Geologic block diagram and cross section depicting a stream environment 45
Figure 3.2 Potential effects of changes in ground-water flow direction on temporal trends in
contaminant concentrations 46
Figure 3.3 Elements of a conceptual site model for monitored natural attenuation of inorganic
contaminants 49
Figure 3.4 Illustration of two approaches for determining attenuation rate constants within a
contaminant plume 50
Figure 3.5 pH-dependent solubility trend of orpiment predicted using two different Gibbs free
energy of formation values 65
VIM
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Tables
Table 1.1 Synopsis of site characterization objective to be addressed throughout the tiered
analysis process and potential supporting data types and/or analysis approaches
associated with each tier 9
Table 1.2 Example software packages for modeling groundwater flow and mass transport 15
Table 1.3 Example software packages for speciation in inorganic geochemical systems 15
Table 1.4 Example software packages for modeling reactive transport in inorganic
geochemical systems 16
Table 1.5 Example internet sources of thermodynamic data useful in constructing
geochemical models 16
Table 1.6 Objectives for performance monitoring of MNA 18
Table 2.1 Important functional groups in humic substances that impact surface charging
behavior and contaminant binding 32
Table 2.2 Major mineral classes in aquifers and soils 34
Table 2.3 Relationships among Q, K, and Q 35
Table 2.4 Range of hydrogen concentrations for a given terminal electron-accepting
process that can be used for classification of the redox status within the
contaminant plume 38
Table 3.1 Sequential extraction procedure of Tessier et al. (1979) 56
Table 3.2 Summary of reagents used to selectively dissolve iron oxides and sulfides 57
Table 3.3 BCR extraction scheme applied to 1 gram of sample 58
Table 3.4 Synopsis of the various surface complexation models (SCMs) commonly
employed to describe solute partitioning to solid surfaces 62
Table 3.5 Ground-water redox parameters and measurement approaches 67
Table 3.6 Methods that may be employed for estimating the oxidation and reduction
capacity for solid materials (from USEPA, 2002) 68
Table 3.7 Standard and emerging techniques for microbial community characterization 70
IX
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Acknowledgments
This document represents a collective work of many individuals with expertise in the policy and techni-
cal 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 their affiliation:
Contributing Author
Richard T Wilkin
Kenneth Lovelace
Stuart Walker
Ronald Wilhelm
Steven Acree
Steve Mangion
Robert W. Puls
Ann Azadpour-Keeley
Robert G. Ford
Patrick V. Brady
James E. Amonette
Paul M. Bertsch
Craig Bethke
Douglas B. Kent
Affiliation
USEPA/ORD, National Risk Management Research Laboratory, Ada, OK
74820
USEPA/OSWER/OSRTI, Washington, DC 20460 (deceased)
USEPA/OSWER/OSRTI, Washington, DC 20460
USEPA/OAR/ORIA, Washington, DC 20460
USEPA/ORD, National Risk Management Research Laboratory, Ada, OK
74820
USEPA/ORD/OSP, Region 1, Boston, MA 02114
USEPA/ORD, National Risk Management Research Laboratory, Ada, OK
74820
USEPA/ORD, National Risk Management Research Laboratory, Ada, OK
74820
USEPA/ORD, National Risk Management Research Laboratory, Cincinnati,
OH 45268
Sandia National Laboratories, Geochemistry Department (MS-0750),
Albuquerque, New Mexico 87185
Pacific Northwest National Laboratory, Fundamental Science Directorate,
Richland, WA 99352
University of Kentucky, College of Agriculture, Lexington, KY 40506
University of Illinois, Department of Geology, Urbana, IL 61801
U.S. Geological Survey, McKelvay Building (MS-465), Menlo Park, CA
94025
Critical and constructive reviews were provided by Jim Weaver (USEPA/ORD National Exposure Research
Laboratory, Athens, GA), George Redden (Idaho National Laboratory, Batelle Energy Alliance), and
Sue Clark (Washington State University, Chemistry Department). 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. This effort is dedicated to the
memory of Kenneth Lovelace, whose insight and patience made it a reality.
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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 context 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 employed, 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 supported 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 ap-
proach to achieve cleanup objectives.
The current document represents the first 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 1, titled "Technical
Basis for Assessment," consists of three sections that describe 1) the conceptual background for natural
attenuation for inorganic contaminants, 2) the technical basis for attenuation of inorganic contaminants
in ground water, and 3) approaches to site characterization to support evaluation of MNA. 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 is-
sues:
1. Demonstration of active contaminant removal from ground water & dissolved plume stability;
2. Determination of the mechanism and rate of attenuation;
3. Determination of the long-term capacity for attenuation and stability of immobilized contami-
nants; and
4. Design of performance monitoring program, including defining triggers for assessing MNA
failure, and establishing a contingency plan.
Detailed discussion is provided on the importance of acquiring site-specific data that define ground-water
hydrogeology and chemistry, the chemical and mineralogical characteristics of aquifer solids, and the
aqueous and solid phase chemical speciation of contaminants within the ground-water plume boundary.
Technical distinctions are drawn between characterization efforts to evaluate the applicability of MNA
as part of a cleanup remedy for organic versus inorganic contaminants. Emphasis is placed on the
need to collect site-specific data supporting evaluation of the long-term stability of immobilized inorganic
contaminants. Also included is discussion on the role of analytical models as one of the tools that may
be employed during the site characterization process. This discussion is intended to provide context to
contaminant-specific site characterization approaches recommended in the remaining two volumes of
this document.
This document is limited to evaluations performed in porous-media settings. Detailed discussion of perfor-
mance 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
xi
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may be incapable of fully delineating 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 con-
taminants may move from source areas and contaminant plumes to impact receptors (e.g., surface water
and indoor air).
Nothing in this document changes Agency policy regarding remedial selection criteria, remedial expec-
tations, or the selection and implementation of MNA. This document does not supersede 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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Section I
Conceptual Background for Monitored Natural Attenuation
Kenneth Lovelace, Stuart Walker, Ronald Wilhelm, Robert Puls, Robert G. Ford, Richard T.
Wilkin, Steven Acree, Steve Mangion, Patrick V. Brady, Craig Bethke
IA. Background and Purpose
IA.1 Document Organization
The purpose of this document is to provide a framework
for assessing the potential application of monitored natural
attenuation as part of the remedy for inorganic contaminant
plumes in ground water. It is organized into three volumes
that provide: Volume 1 - a general overview of the framework
and technical requirements for application of Monitored
Natural Attenuation (MNA); Volume 2 - contaminant-specific
discussions addressing potential attenuation processes and
site characterization requirements for non-radionuclides,
and Volume 3 - contaminant-specific discussions address-
ing potential attenuation processes and site characterization
requirements for radionuclides. Volume 1 is divided into
three sections that address the regulatory and conceptual
background for natural attenuation, the technical basis for
natural attenuation of inorganic contaminants, and site
characterization approaches to support assessment and
application of MNA. The contaminant-specific chapters
in Volumes 2 and 3 provide an overview of contaminant
geochemistry, applicable natural attenuation processes,
and specific site characterization requirements. Criteria for
selecting specific contaminants for these detailed overviews
are described below.
The non-radionuclide contaminants selected for this docu-
ment include: arsenic (As), cadmium (Cd), chromium (Cr),
copper (Cu), lead (Pb), nickel (Ni), nitrate, perchlorate, and
selenium (Se). The selection of these contaminants by
USEPA was based on several criteria. First, a 1994 booklet
containing information regarding common chemicals found
at Superfund sites throughout the nation was consulted
(USEPA, 1994). The most commonly found inorganic con-
taminants were included for consideration in this document.
Another document specific to metal-contaminated Super-
fund sites (USEPA, 1995) identified arsenic (As), cadmium
(Cd), chromium (Cr), and lead (Pb) as primary contaminants
of concern based on toxicity, industrial use, and frequency
of occurrence at Superfund sites. Second, selection was
based on chemical behavior considering chemical traits
such as: toxicity, ion charge (cation vs. anion), transport
behavior (conservative vs. non-conservative), and redox
chemistry to cover a broad range of geochemical behavior
(USEPA,1999a; USEPA, 1999b; USEPA, 2004). Finally,
USEPA regional staff were asked to nominate inorganic
contaminants that occurred frequently or that were prob-
lematic in their Regions. The above list of nine inorganic
contaminants was selected from this process.
The radionuclide contaminants selected for this document
include: americium (Am), cesium (Cs), iodine (I), neptunium
(Np), plutonium (Pu), radium (Ra), radon (Rn), technetium
Tc), thorium (Th), tritium, strontium (Sr), and uranium (U).
The selection of these contaminants by EPA was based
on two criteria. First, a selected element had to be one
of high priority to the site remediation or risk assessment
activities of the USEPA (USEPA, 1993; USEPA, 2002).
Second, selection was based on chemical behavior con-
sidering chemical traits such as: toxicity, cations, anions,
conservatively transported, non-conservatively transported,
and redox sensitive elements (USEPA, 1999b; USEPA,
2004). By using these characteristics of the contami-
nants, the general geochemical behavior of a wide range
of radionuclide contaminants could be covered as well as
the chemical classes that make up the Periodic Table. In
addition, this selection accounts for many daughter and
fission product contaminants that result from radioactive
decay. This is important as the decay of radioisotopes can
produce daughter products that may differ both physically
and chemically from their parents. The selection of radio-
nuclide contaminants for this document is representative
of these characteristics.
IA.2 Purpose of Document
This document is intended to provide a technical resource
for determining whether MNA is likely to be an effective
remedial approach for inorganic contaminants1 in ground
water. This document is intended to be used during the
remedial investigation and feasibility study phases of a
Superfund cleanup, or during the equivalent phases of a
RCRA Corrective Action (facility investigation and corrective
measures study, respectively). The decision to select MNA
as the remedy (or part of the remedy) will be made in a
Superfund Record of Decision (ROD) or a RCRA Statement
of Basis (or RCRA permit).
The USEPA expects that users of this document will include
USEPA and State cleanup programs and their contractors,
especially those individuals responsible for evaluating al-
ternative cleanup methods for a given site or facility. The
overall policy for use of MNA in OSWER cleanup programs
is described in the April 21, 1999 OSWER Directive titled,
"Use of Monitored Natural Attenuation at Superfund, RCRA
Corrective Action and Underground Storage Tank Sites"
(Directive No. 9200.4-17P).
The term "inorganic contaminants" is used in this document as a
generic term for metals and metalloids (such as arsenic); and also
refers to radiologic as well as non-radiologic isotopes.
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Both radiological and non-radiological inorganic contami-
nants are discussed in this document. There are two rea-
sons for this. First, except for radioactive decay, the potential
attenuation processes affecting inorganic contaminants
are the same for both contaminant types. Second, several
OSWER directives clarify the USEPA's expectation that the
decision-making approach and cleanup requirements used
at CERCLA sites will be the same for sites with radiological
and non-radiological inorganic contaminants, except where
necessary to account for the technical differences between
the two types of contaminants. Also, the 1999 OSWER
Directive specified that the decision process for evaluat-
ing MNA as a potential remediation method should be the
same for all OSWER cleanup programs.
This document is intended to provide an approach for
evaluating MNA as a possible cleanup method for contami-
nated ground water. Although the focus of the document
is on ground water, the unsaturated zone is discussed as
a source of contaminants to ground water. Emphasis is
placed on developing a more complete evaluation of the
site through development of a conceptual site model2 based
on an understanding of the attenuation mechanisms, the
geochemical conditions governing these mechanisms, the
capacity of the aquifer to sustain attenuation of the contami-
nant mass and prevent future contaminant migration, and
indicators that can be used to monitor MNA performance.
This document focuses on technical issues and is not in-
tended to address policy considerations or specific regula-
tory or statutory requirements. The USEPA expects that this
document will be used in conjunction with the 1999 OSWER
Directive (USEPA, 1999c). Users of this document should
realize that different Federal and State remedial programs
may have somewhat different remedial objectives. For ex-
ample, the CERCLA and RCRA Corrective Action programs
generally require that remedial actions: 1) prevent exposure
to contaminated ground water, above acceptable risk levels;
2) minimize further migration of the plume; 3) minimize
further migration of contaminants from source materials;
and 4) restore ground-water conditions to cleanup levels
appropriate for current or future beneficial uses, to the
extent practicable. Achieving such objectives could often
require that MNA be used in conjunction with other "active"
remedial methods. For other cleanup programs, remedial
objectives may be focused on preventing exposures above
acceptable levels. Therefore, it is imperative that users of
this document be aware of and understand the Federal
and State statutory and regulatory requirements, as well as
policy considerations that apply to a specific site for which
this document will be used to evaluate MNA as a remedial
option. As a general practice, individuals responsible for
evaluating remedial alternatives should check with the over-
2 A conceptual site model is a three-dimensional representation that
conveys what is known or suspected about contamination sources,
release mechanisms, and the transport and fate of those contami-
nants. The conceptual model provides the basis for assessing poten-
tial remedial technologies at the site. "Conceptual site model" is not
synonymous with "computer model"; however, a computer model may
be helpful for understanding and visualizing current site conditions or
for predictive simulations of potential future conditions.
seeing regulatory agency to identify likely characterization
and cleanup objectives for a particular site prior to investing
significant resources.
Use of this document is generally inappropriate in complex
fractured bedrock or karst aquifers. In these situations the
direction of ground water flow can not be predicted directly
from the hydraulic gradient, and existing techniques may
not be capable of identifying the pathway along which
contaminated groundwater moves through the subsurface.
Understanding the contaminant flow field in the subsurface
is essential for a technically justified evaluation of an MNA
remedial option. MNA will not generally be appropriate
where site complexities preclude adequate monitoring
(USEPA, 1999c).
Because documentation of natural attenuation requires
detailed site characterization, the data collected can be
used to compare the relative effectiveness of other remedial
options and natural attenuation. The technical information
contained in this document can be used as a point of refer-
ence to evaluate whether MNA by itself, or in conjunction
with other remedial technologies, is sufficient to achieve
site-specific remedial objectives.
IA.3 Applicable Regulatory Criteria
All remedial actions at CERCLA sites must be protective
of human health and the environment and comply with ap-
plicable or relevant and appropriate requirements (ARARs)
unless a waiver is justified. Cleanup levels for response ac-
tions under CERCLA are developed based on site-specific
risk assessments, ARARs, and/or to-be-considered material
(TBCs). The determination of whether a requirement is
applicable, or relevant and appropriate, must be made on
a site-specific basis (see 40 CFR §300.400(g)).
"EPA expects to return usable ground waters to their
beneficial uses whenever practicable" (see 40 CFR §30
0.430(a)(1)(iii)(F)). In general, drinking water standards
provide relevant and appropriate cleanup levels for ground
waters that are a current or potential source of drinking
water. However, drinking water standards generally are
not relevant and appropriate for ground waters that are
not a current or potential source of drinking water (see
55 FR 8732, March 8, 1990). Drinking water standards
include federal maximum contaminant levels (MCLs) and/
or non-zero maximum contaminant level goals (MCLGs)
established under the Safe Drinking Water Act (SDWA),
or more stringent state drinking water standards. Other
regulations may also be ARARs as provided in CERCLA
§121(d)(2)(B).
IA.4 Policy Framework for Use of MNA
The term "monitored natural attenuation" is used in this
document when referring to a particular approach to re-
mediation. MNA is defined in the 1999 OSWER Directive
as follows:
"...the reliance on natural attenuation processes
(within the context of a carefully controlled and
monitored site cleanup approach) to achieve site-
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specific remediation objectives within a time frame
that is reasonable compared to that offered by
other more active methods. The 'natural attenuation
processes' that are at work in such a remediation
approach include a variety of physical, chemical,
or biological processes that, under favorable condi-
tions, act without human intervention to reduce the
mass, toxicity, mobility, volume, or concentration of
contaminants in soil or groundwater. These in-situ
processes include biodegradation; dispersion; dilu-
tion; sorption; volatilization; radioactive decay; and
chemical or biological stabilization, transformation,
or destruction of contaminants. (USEPA, 1999c,
page 3.)
Even though several physical, chemical, and biological
processes are included in the above definition, the 1999
OSWER Directive goes on to state a preference for those
processes that permanently degrade or destroy contami-
nants, and for use of MNA for stable or shrinking plumes,
as noted below:
"When relying on natural attenuation processes
for site remediation, EPA prefers those processes
that degrade or destroy contaminants. Also, EPA
generally expects that MNA will only be appropriate
for sites that have a low potential for contaminant
migration." (USEPA, 1999c, page 3.)
"MNA should not be used where such an approach
would result in either plume migration or impacts to
environmental resources that would be unaccept-
able to the overseeing regulatory authority. There-
fore, sites where the contaminant plumes are
no longer increasing in extent, or are shrinking,
would be the most appropriate candidates for
MNA remedies." (USEPA, 1999c, page 18.)
Control of contaminant sources is also an important as-
pect of EPA's policy. The actual policy language is given
below:
"Control of source materials is the most effective
means of ensuring the timely attainment of reme-
diation objectives. EPA, therefore, expects that
source control measures will be evaluated for
all contaminated sites and that source control
measures will be taken at most sites where
practicable. At many sites it will be appropriate
to implement source control measures during the
initial stages of site remediation ("phased remedial
approach"), while collecting additional data to de-
termine the most appropriate groundwater remedy."
(USEPA, 1999c, page 22.)
The 1999 OSWER Directive also provides a few general
guidelines for use of MNA as a remedial approach for
inorganic contaminants. The key policy concerns are that
the specific mechanisms responsible for attenuation of in-
organic contaminants should be known at a particular site,
and the stability of the process should be evaluated and
shown to be protective under anticipated changes in site
conditions. The actual policy language is given below:
MNA may, under certain conditions (e.g., through
sorption or oxidation-reduction reactions), effec-
tively reduce the dissolved concentrations and/or
toxic forms of inorganic contaminants in groundwa-
ter and soil. Both metals and non-metals (includ-
ing radionuclides) may be attenuated by sorption3
reactions such as precipitation, adsorption on
the surfaces of soil minerals, absorption into the
matrix of soil minerals, or partitioning into organic
matter. Oxidation-reduction (redox) reactions can
transform the valence states of some inorganic
contaminants to less soluble and thus less mobile
forms (e.g., hexavalent uranium to tetravalent ura-
nium) and/or to less toxic forms (e.g., hexavalent
chromium to trivalent chromium). Sorption and
redox reactions are the dominant mechanisms
responsible for the reduction of mobility, toxicity,
or bioavailability of inorganic contaminants. It is
necessary to know what specific mechanism (type
of sorption or redox reaction) is responsible for the
attenuation of inorganics so that the stability of
the mechanism can be evaluated. For example,
precipitation reactions and absorption into a soil's
solid structure (e.g., cesium into specific clay
minerals) are generally stable, whereas surface
adsorption (e.g., uranium on iron-oxide minerals)
and organic partitioning (complexation reactions)
are more reversible. Complexation of metals or
radionuclides with carrier (chelating) agents (e.g.,
trivalent chromium with EDTA) may increase their
concentrations in water and thus enhance their
mobility. Changes in a contaminant's concentra-
tion, pH, redox potential, and chemical speciation
may reduce a contaminant's stability at a site and
release it into the environment. Determining the
existence, and demonstrating the irreversibility of
these mechanisms is important to show that a MNA
remedy is sufficiently protective.
In addition to sorption and redox reactions, radio-
nuclides exhibit radioactive decay and, for some,
a parent-daughter radioactive decay series. For
example, the dominant attenuating mechanism
of tritium (a radioactive isotopic form of hydrogen
with a short half-life) is radioactive decay rather
than sorption. Although tritium does not generate
radioactive daughter products, those generated by
some radionulides (e.g., Am-241 and Np-237 from
Pu-241) maybe more toxic, have longer half-lives,
and/or be more mobile than the parent in the decay
series. Also, it is important that the near surface or
3 When a contaminant is associated with a solid phase, it is usually not
known if the contaminant is precipitated as a three-dimensional mo-
lecular coating on the surface of the solid, adsorbed onto the surface
of the solid, absorbed into the structure of the solid, or partitioned into
organic matter. "Sorption" will be used in this Directive to describe, in
a generic sense (i.e., without regard to the precise mechanism) the
partitioning of aqueous phase constituents to a solid phase.
-------
surface soil pathways be carefully evaluated and
eliminated as potential sources of external direct
radiation exposure.4
Inorganic contaminants persist in the subsurface
because, except for radioactive decay, they are
not degraded by the other natural attenuation pro-
cesses. Often, however, they may exist in forms that
have low mobility, toxicity orbioavailabilitysuch that
they pose a relatively low level of risk. Therefore,
natural attenuation of inorganic contaminants is
most applicable to sites where immobilization or ra-
dioactive decay is demonstrated to be in effect and
the process/mechanism is irreversible. (USEPA,
1999c, pages 8-9.)
The 1999 OSWER Directive provides the context for the
Agency's expectations for evaluating the feasibility of em-
ploying MNA as part of a cleanup remedy for contaminated
ground water. As indicated by the sections from the Direc-
tive that are transcribed above, it also points out specific
issues concerning what constitutes natural attenuation for
inorganic contaminants. In practice, most of the techni-
cal experience developed to date has primarily dealt with
evaluations of MNA as applied to remediation of organic
contaminant plumes. While this experience provides some
perspective for the scope of site characterization that may
be warranted to evaluate MNA for inorganic contaminants,
there are some important distinctions that bear on the
types of required data and the approaches available to
obtain these data. The following section elaborates these
distinctions in order to provide context for the technical
aspects relevant to MNA for inorganic contaminants and
the steps needed to implement a technically defensible site
characterization effort.
IB. Relevant Distinctions in Site
Characterization for MNA of Inorganic
Contaminants
As stated within the OSWER Directive on MNA (USEPA,
1999c), natural attenuation processes are those that 're-
duce mass, toxicity, mobility, volume or concentration of
contaminants'. Inorganic contaminants discussed within
this document include both non-radioactive and radioac-
tive constituents. For radioactive contaminants, radioactive
decay processes result in the reduction of risk derived
from radiation exposure. The rates of radioactive decay
(characterized by the decay half-life) are known for the
radioisotopes of concern, thus facilitating this aspect of site
characterization. Guidelines for assessing the feasibility of
MNA as a component of ground-water cleanup for radio-
4 External direct radiation exposure refers to the penetrating radiation
(i.e., primarily gamma radiation and x-rays) that may be an important
exposure pathway for certain radionuclides in near surface soils. Un-
like chemicals, radionuclides can have deleterious effects on humans
without being taken into or brought in contact with the body due to
high-energy particles emitted from near surface soils. Even though
the radionuclides that emit penetrating radiation may be immobilized
due to sorption or redox reactions, the resulting contaminated near
surface soil may not be a candidate for a MNA remedy as a result of
this exposure risk.
nuclides are provided in Volume 3 of this document. For
non-radioactive inorganic contaminants and radionuclides
possessing long decay half-lives, immobilization within the
aquifer via sorption to aquifer solids provides the primary
means for attenuation of the ground-water plume. In gen-
eral, an inorganic contaminant can be transferred between
solid, liquid, or gaseous phases present within the aquifer,
but the contaminant will always be present. Contaminant
immobilization will prevent transport to sensitive receptors
at points of compliance. There are limited examples where
degradation of inorganic contaminants may be a viable
attenuation process (e.g., biological degradation of nitrate
or perchlorate), but degradation is not a viable process for
most of the inorganic contaminants discussed in this docu-
ment. For inorganic contaminants subject to degradation
or reductive transformation processes, the supporting site
characterization will likely be consistent with the approach
employed to assess MNA for organic contaminant plumes
(e.g., USEPA, 1998; USEPA, 2001; see also specific discus-
sions for nitrate and perchlorate in Volume 2). The following
discussion provides context for the potential significance
of immobilization as a means for natural attenuation of
inorganic contaminants in ground water.
There is an important distinction between site character-
ization as applied to assessment of MNA for organic and
inorganic contaminants. For organic contaminants, site
characterization typically is focused towards determining the
mechanism of contaminant degradation and the capacity
of site conditions to sustain degradation for treatment of
the mass of contaminant within the plume. This analysis
may include identification of ground-water characteristics
and degradation byproducts that are characteristic for con-
taminant degradation. Thus, much of the emphasis on site
characterization for MNA of organic contaminants has been
directed towards the collection and analysis of ground-water
samples. In some cases, this characterization effort may
have been supplemented with the analysis of contaminant
degradation behavior through the use of microcosm experi-
ments employing aquifer solids collected within the plume
boundary. For inorganic contaminants in which immobili-
zation onto aquifer solids provides the primary means for
attenuation of the ground-water plume, characterization of
the solid substrate within the aquifer plays a more significant
role during site assessment. In this case risk reduction in
ground water is realized through the sorption of the inorgan-
ic contaminant onto aquifer solids in combination with the
long-term stability of the immobilized contaminant to resist
remobilization due to changes in ground-water chemistry.
The importance of this distinction between natural attenua-
tion for organic and inorganic contaminants is emphasized
in Figure 1.1. In essence, for inorganic contaminants one
can consider the existence of two distinct 'plumes' within
the boundary of the ground-water plume: 1) the dissolved
or "mobile" plume (including dissolved contaminant and
contaminant associated with mobile colloids), and 2) the
solid phase or "immobile" plume resulting from sorption of
the contaminant to aquifer solids (Figure 1.1). Thus, for
inorganic contaminants there are two overriding objectives
to address through site characterization:
-------
1) Demonstration of removal of the inorganic contami-
nant from the dissolved phase leading to a stable
or shrinking ground-water plume and,
2) Demonstration of stabilization of the inorganic
contaminant immobilized onto aquifer solids such
that future re-mobilization will not occur to a level
that threatens health of environmental receptors.
Evaluating the overall success of natural attenuation for
inorganic contaminant remediation will require demonstrat-
ing that the rate and capacity for inorganic contaminant
attenuation meets regulatory objectives and, in addition,
that inorganic contaminant immobilization is sustainable to
the extent that future health risks are eliminated. The latter
requirement necessitates identifying the chemical specia-
tion of the inorganic contaminant partitioned to the solid
phase. This information is critical towards identifying the
process controlling attenuation and evaluating the long-term
stability of the immobilized contaminant relative to observed
or anticipated changes in ground-water chemistry.
Site characterization to support evaluation of MNA as a
remedial alternative will involve assessment of contaminant
transport in the aquifer. In general terms, this process will
include assessment of ground-water hydrology and the
biogeochemical processes that control contaminant migra-
tion within the plume. Defining the processes that control
contaminant immobilization (or degradation) along the paths
of ground-water flow will necessitate collection of a range
of data that define the dynamics of system hydrology, the
chemical characteristics of ground water, and the proper-
ties of the aquifer solids. In order to screen out sites that
Organic
Time
are inappropriate for selection of MNA, it is recommended
that collection of site-specific data be conducted in stages
that serve to minimize expenditures while providing insight
into the potential existence of natural processes that may
attenuate contaminant migration. Description of a tiered
analysis approach for organizing site characterization tasks
is provided in the following section.
1C. Tiered Analysis Approach to Site
Characterization
Site characterization to support evaluation and selection of
MNA as part of a cleanup action for inorganic contaminant
plumes in ground water will involve a detailed analysis of site
characteristics controlling and sustaining attenuation. The
level of detailed data that may be required to adequately
characterize the capacity and stability of natural processes
to sustain plume attenuation will likely necessitate signifi-
cant resource outlays. Thus, it is recommended that site
characterization be approached in a step-wise manner
to facilitate collection of data necessary to progressively
evaluate the existing and long-term effectiveness of natural
attenuation processes within the aquifer. Implementation
of a tiered analysis approach provides an effective way to
screen sites for MNA that is cost effective because it priori-
tizes and limits the data that is needed for decision making
at each screening step. Conceptually a tiered analysis
approach seeks to progressively reduce uncertainty as
site-specific data are collected. The decision-making ap-
proach presented in this document includes three decision
tiers that require progressively greater information on which
to assess the likely effectiveness of MNA as a remedy for
Inorganic
Original Plume
Boundary
Mobile Contaminant
^ssrm
— ^ Immobile Contaminant
Mobile plume
shrinkage due to
degradation or
immobilization onto
aquifer sediments
Immobilized
inorganic
contaminant still
present on
aquifer solids
Figure 1.1 Conceptual distinction between organic versus inorganic contaminant plume behavior where natural
processes are active within the ground-water aquifer. Natural attenuation of inorganic contaminants is
viable only if the immobilized contaminant remains stable and resistant to remobilization during changes
in ground-water chemistry.
-------
inorganic contaminants in ground water. The fourth tier is
included to emphasize the importance of determining ap-
propriate parameters for long-term performance monitoring,
once MNA has been selected as part of the remedy. Data
collection and evaluation within the tiered analysis approach
would be structured as follows:
I. Demonstration that the ground-water plume is not
expanding and that sorption of the contaminant
onto aquifer solids is occurring where immobiliza-
tion is the predominant attenuation process;
II. Determination of the mechanism and rate of the
attenuation process;
III. Determination of the capacity of the aquifer to at-
tenuate the mass of contaminant within the plume
and the stability of the immobilized contaminant to
resist re-mobilization, and;
IV. Design performance monitoring program based on
the mechanistic understanding developed for the
attenuation process, and establish a contingency
plan tailored to site-specific characteristics.
Elaboration on the objectives to be addressed and the types
of site-specific data to be collected under each successive
tier is provided below.
IC.1 Tier I
The objective under Tier I analysis would be to eliminate
sites where site characterization indicates that the ground-
water plume is continuing to expand in aerial or vertical
extent. For contaminants in which sorption onto aquifer
solids is the most feasible attenuation process, an additional
objective would be to demonstrate contaminant uptake onto
aquifer solids. Analysis of ground-water plume behavior
at this stage is predicated on adequate aerial and vertical
delineation of the plume boundaries. Characterization of
ground-water plume expansion could then be supported
through analysis of current and historical data collected
from monitoring wells installed along the path of ground-
water flow. An increasing temporal trend in contaminant
concentration in ground-water at monitoring locations down
gradient from a source area is indicative that attenuation
is not occurring sufficient to prevent ground-water plume
expansion. Determination of contaminant sorption onto
aquifer solids could be supported through the collection of
aquifer cores coincident with the locations of ground-water
data collection and analysis of contaminant concentrations
on the retrieved aquifer solids. Illustration of the type of
data trend anticipated for a site where sorption actively
attenuates contaminant transport is provided in Figure 1.2.
The spatial distribution in aqueous and solid contaminant
Aquifer Core Samples
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Figure 1.2 Conceptual depiction of the data collection effort to demonstrate whether sorption to aquifer solids atten-
uates contaminant transport in ground water. The left side of the diagram provides a cross-sectional view
of the spatial distribution of the contaminant concentration in ground water and co-located aquifer solids
for a site where sorption attenuates contaminant transport. The trend in aqueous and solid contaminant
concentrations for this scenario is depicted in Panel (A) to the right. Panel (B) depicts the relationship
between aqueous and solid contaminant concentrations for a site where sorption does not attenuate
contaminant transport.
-------
concentrations for a site where sorption attenuates
contaminant migration is depicted on the left side of the
illustration. Anticipated relationships between aqueous
and solid contaminant concentrations for sites with and
without active contaminant attenuation via sorption are
depicted on the right side of the illustration in Panels (A)
and (B), respectively. Specifically, where sorption onto
aquifer solids is occurring, there should be an increasing
trend in solid phase contaminant concentrations as a
function of increasing aqueous concentration. In contrast,
no change in solid phase contaminant concentrations as a
function of increasing aqueous concentration is indicative
that attenuation is not occurring. Ultimately, sites that
demonstrate ground-water plume expansion and a lack of
contaminant sorption (for contaminants subject to sorption)
would be eliminated from further consideration of MNA as
part of the cleanup remedy.
IC.2 Tier II
The objective under Tier II analysis would be to eliminate
sites where further analysis shows that attenuation rates
are insufficient for attaining cleanup objectives established
for the site within a timef rame that is reasonable compared
to other remedial alternatives, (see USEPA, 1999c, pages
19-21, for a discussion of "reasonable timef rame for reme-
diation".) Data collection and analysis performed for Tier
II would indicate whether MNA processes are capable of
achieving remediation objectives, based on current geo-
chemical conditions at the site. This data collection effort
would also be designed to support identification of the spe-
cific mechanism(s) controlling contaminant attenuation.
An estimate of attenuation rates for inorganic contaminants
will typically involve calculation of the apparent transfer
of mass from the aqueous to the solid phase, based
on sampling of ground water and/or aquifer solids. It is
recommended that these estimates be based as much
as possible on field measurements rather than model-
ing predictions. A recommended approach is to identify
hydrostratigraphic units for the site and develop a ground
water flow model which can be used to estimate ground
water seepage velocities in each of these units (Further
information on ground water flow models is provided in
Section I.D.) These seepage velocities can be combined
with measured contaminant concentrations to estimate
mass flux (mass per time per area) for each contaminant,
in each hydrostratigraphic unit. The necessary data might
include physical parameters such as hydraulic conductivi-
ties within the aquifer and hydraulic gradients. Changes
in mass flux can then be used to estimate mass loss from
the aqueous phase since the last sampling event, which
is assumed to be the apparent attenuation rate. (Further
information on estimating attenuation rates is provided in
Section IIIA.5.)
Determination of attenuation mechanism will depend on
collection of data to define ground-water chemistry, aqui-
fer solids composition and mineralogy, and the chemical
speciation of the contaminant in ground water and as-
sociated aquifer solids. This will entail a significant effort
in the site-specific data collection effort, but provides the
underpinning for further evaluation of the performance of
MNA to be addressed in subsequent stages of the site
characterization process. The goal of this characterization
effort is to identify the aqueous and solid phase constituents
within the aquifer that control contaminant attenuation. This
data collection effort may include collection of field water
quality data (e.g., pH, dissolved oxygen, alkalinity, ferrous
iron, and dissolved sulfide), laboratory measurements of
ground-water and aquifer solids chemical composition,
microbial characteristics and/or mineralogy of the aquifer
solids (as relevant to degradation or immobilization), and
the chemical speciation of the contaminant in ground-water
and/or the aquifer solids. Contaminant speciation refers to
both oxidation state characterizations [e.g., As(lll) vs. As(V)]
as well as specific associations with chemical constituents
in aquifer solids (e.g., precipitation of Pb carbonate vs.
adsorption of Pb to iron oxides). Evaluations of the sub-
surface microbiology may be necessary in situations where
biotic processes play a direct or indirect role in governing
contaminant attenuation. Indirect microbial influence on
contaminant attenuation includes situations in which the
predominant characteristics of the ground-water chemistry
are controlled by microbial oxidation-reduction reactions.
This situation may be more predominant in plumes in
which readily degradable organic contaminants, such as
hydrocarbons or chlorinated solvents, are also present.
Ultimately, mechanistic knowledge of the attenuation pro-
cess along with a detailed knowledge of the ground-water
flow field provides the basis for subsequent evaluations
to assess the long-term capacity of the aquifer to sustain
contaminant attenuation.
/C.3 Tier III
The objective under Tier III would be to eliminate sites
where site data and analysis show that there is insufficient
capacity in the aquifer to attenuate the contaminant mass
to ground-water concentrations that meet regulatory objec-
tives or that the stability of the immobilized contaminant is
insufficient to prevent re-mobilization due to future changes
in ground-water chemistry. Possible factors that could result
in an insufficient capacity for attenuation include:
1. changes in ground-water chemistry result in slower
rates of attenuation,
2. insufficient mass flux of aqueous constituents that
participate in the attenuation reaction, and/or
3. insufficient mass of solid constituents in aquifer
solids that participate in the attenuation reaction.
These factors pertain to situations where either degrada-
tion or immobilization is the primary attenuation process.
For immobilized contaminants, factors to consider relative
to the long-term stability of the attenuated contaminant in-
clude changes in ground-water chemistry that could result
in release of the contaminant from aquifer solids due to
desorption from solid surfaces or dissolution of precipitates.
For example, contaminant desorption could be caused by
changes in ground-water pH, since the degree of adsorp-
tion is typically sensitive to this parameter. Alternatively,
dissolution of a contaminant attenuated as a carbonate
-------
precipitate may result from decreases in ground-water pH
and alkalinity.
Assessment of attenuation capacity will depend on knowl-
edge of the flux of contaminants and associated reactants
in ground-water, as well as the mass distribution of reac-
tive aquifer solids along ground-water flow paths. In order
to conduct this type of evaluation, adequate information
is needed on the heterogeneity of the ground-water flow
field, and the spatial and/or temporal variability in the dis-
tribution of aqueous and solids reactants within the plume.
For situations where ground-water chemistry is governed
by microbial processes, seasonal variations may exert an
indirect influence on the effective capacity within the aquifer
at any point in time. The general approach that can be taken
is to estimate the attenuation capacity within the plume
boundaries and compare this capacity with the estimated
mass flux of aqueous phase contaminants emanating
from source areas based on site-specific data. Exploring
alternatives to minimize contaminant release from source
areas may prove beneficial for sites that possess insufficient
capacity to adequately attenuate the ground-water plume.
Ultimately, this points to the critical importance of a detailed
characterization of the system hydrology.
Assessment of the stability of an immobilized contaminant
can be evaluated through a combination of laboratory
testing and chemical reaction modeling within the context
of existing and anticipated site conditions. Both analysis
approaches can be developed based on the information
gathered during Tier II efforts to characterize the specific
attenuation process active within the ground-water plume.
Through Tier II analysis, a specific attenuation reaction
was defined that identified critical reaction parameters such
as the identity of dissolved constituents that participated
in the process. In addition, mechanistic understanding of
the overall reaction provides the context for evaluating site
conditions or dissolved constituents that may interfere with
or reduce the efficiency of the attenuation reaction. For ex-
ample, sites where the contaminant plume is reducing (e.g.,
sulfate-reducing conditions) while ambient ground-water is
oxidizing may be susceptible to future influxes of dissolved
oxygen. In this situation, the attenuation process may be
due to precipitation of sulfides under sulfate-reducing condi-
tions within the plume. Future exposure of these sulfides
to oxygen may result in dissolution of the sulfide precipitate
along with release of the contaminant back into ground wa-
ter. Alternatively, sites where attenuation is predominated
by contaminant adsorption onto existing aquifer solids may
be sensitive to future influx of dissolved constituents due to
land use changes that alter either the source or chemical
composition of ground-water recharge. The sensitivity to
contaminant re-mobilization can be assessed via labora-
tory tests employing aquifer solids collected from within the
plume boundaries that can be exposed to solutions that
mimic anticipated ground-water chemistries (e.g., ambient
ground-water samples or synthetic solutions in which the
concentrations of specific dissolved constituents can be
systematically varied). A supplementary avenue to test
contaminant stability could include use of chemical reac-
tion models with adequate parameterization to replicate
both the attenuation reaction as well as changes in water
composition that may interfere with attenuation. The util-
ity of this type of modeling analysis would be the ability to
efficiently explore contaminant solubility under a range of
hypothetical ground-water conditions in order to identify the
ground-water parameters to which the attenuation reaction
may be most sensitive.
It is feasible to consider implementation of MNA as a
component of the ground-water remedy if the analysis con-
ducted through the previous Tiers indicates that the aquifer
within the plume boundaries supports natural attenuation
processes with sufficient efficiency, capacity, and stability.
The technical knowledge obtained through identification of
the specific attenuation mechanism and the sensitivity of the
attenuation process to changes in ground-water chemistry
can then be employed in designing a monitoring program
that tracks continued performance of the MNA remedy.
1C A Tier IV
The objective under Tier IV analysis is to develop a monitor-
ing program to assess long-term performance of the MNA
remedy and identify alternative remedies that could be
implemented for situations where changes in site conditions
could lead to remedy failure. Site data collected during
characterization of the attenuation process will serve to
focus identification of alternative remedies that best match
site-specific conditions. The monitoring program will consist
of establishing a network of wells: 1) that provide adequate
aerial and vertical coverage to verify that the ground-water
plume remains static or shrinks, and 2) that provide the
ability to monitor ground-water chemistry throughout the
zones where contaminant attenuation is occurring. It is
recommended that the performance monitoring program
include assessment of the consistency in ground-water
flow behavior, so that adjustments to the monitoring net-
work could be made to evaluate the influence of potential
changes in the patterns of ground-water recharge to or
predominant flow direction within the plume. In addition
to monitoring ground-water parameters that track the at-
tenuation reaction, periodic monitoring of parameters that
track non-beneficial changes in ground-water conditions is
also recommended. Monitoring the attenuation reaction
will include continued verification of contaminant removal
from ground water, but will also include tracking trends in
other reactants that participate in the attenuation reaction
(possible examples include pH, alkalinity, ferrous iron, and
sulfate). For sites in which contaminant immobilization
is the primary attenuation process, periodic collection of
aquifer solids may be warranted to verify consistency in
reaction mechanism. It is recommended that the selection
of ground-water parameters to be monitored also include
constituents that provide information on continued stability
of the solid phase with which an immobilized contaminant
is associated. Examples of this type of parameter might
include ferrous iron or sulfate to track dissolution of iron ox-
ides or sulfide precipitates, respectively. Non-contaminant
performance parameters such as these will likely serve as
"triggers" to alert site managers to potential remedy failure
or performance losses, since the attenuation reaction will
-------
respond to these changed conditions. Since increases in
mobile contaminant concentrations may be delayed relative
to changes in site conditions, these monitoring parameters
may improve the ability of site managers to evaluate and
address the potential for ground-water plume expansion.
In summary, the tiered analysis process provides a means
to organize the data collection effort in a cost-effective
manner that allows the ability to eliminate sites at interme-
diate stages of the site characterization effort. A general
synopsis of the objectives along with possible analysis
approaches and/or data types to be collected under each
tier is provided in Table 1.1. The types of data collected
early in the site characterization process would typically be
required for selection of appropriate engineered remedies,
including characterization of the system hydrology, ground-
water chemistry, contaminant distribution, and the aqueous
speciation of the contaminant. These system characteristics
can have direct influence on the selection of pump-and-treat
or in-situ remedies best suited to achieve cleanup objec-
tives for inorganic contaminants. This limits any loss on
investment in site characterization for sites where selection
of MNA as part of the ground-water remedy is ultimately
determined not viable. The primary objective of progressing
through the proposed tiered site analysis steps is to reduce
uncertainty in the MNA remedy selection.
The remaining discussion in this section of Volume 1 will
elaborate on two issues that have been introduced above,
specifically the use of models in site characterization
and general factors to consider for implementation of a
long-term performance monitoring program. These topics
are addressed at this juncture to allow greater focus to
discussions later in this volume pertaining specifically to
Table 1.1 Synopsis of site characterization objective to be addressed throughout the tiered analysis process and
potential supporting data types and/or analysis approaches associated with each tier.
Tier
Objective
Potential Data Types and Analysis
Demonstrate active con-
taminant removal from
ground water
Ground-water flow direction (calculation of hydraulic gradients); aquifer
hydrostratigraphy
Contaminant concentrations in ground water and aquifer solids
General ground-water chemistry data for preliminary evaluation of con-
taminant degradation
Determine mechanism and
rate of attenuation
Detailed characterization of system hydrology (spatial and temporal
heterogeneity; flow model development)
Detailed characterization of ground-water chemistry
Subsurface mineralogy and/or microbiology
Contaminant speciation (ground water & aquifer solids)
Evaluate reaction mechanism (site data, laboratory testing, develop
chemical reaction model)
Determine system capacity
and stability of attenu-
ation
Determine contaminant & dissolved reactant fluxes (concentration data
& water flux determinations)
Determine mass of available solid phase reactant(s)
Laboratory testing of immobilized contaminant stability (ambient ground
water; synthetic solutions)
Perform model analyses to characterize aquifer capacity and to test
immobilized contaminant stability (hand calculations, chemical reaction
models, reaction-transport models)
IV
Design performance
monitoring program
and identify alternative
remedy
Select monitoring locations and frequency consistent with site heteroge-
neity
Select monitoring parameters to assess consistency in hydrology, at-
tenuation efficiency, and attenuation mechanism
Select monitored conditions that "trigger" re-evaluation of adequacy of
monitoring program (frequency, locations, data types)
Select alternative remedy best suited for site-specific conditions
-------
attenuation processes (Volume 1, Section II) and the types
of site characterization data needed for their identification
(Volume 1, Section III). The following discussion provides
perspective on the role of model applications in the site
characterization process, the types of models that might
be employed to help meet the objectives set forth under
each tier, and potential limitations in the availability and
adequacy of available model codes.
ID. Role of Modeling in the Tiered Analysis
Approach
Design of the site characterization effort and analysis of
site-specific data in support of assessing the suitability of
MNA as a component of the ground-water remedy is de-
pendent on development of a Conceptual Site Model (CSM)
that identifies site conditions and processes that influence
contaminant transport. The CSM also provides the under-
pinning for selecting and developing model applications that
provide a set of tools for evaluating transport processes,
reaction mechanisms, attenuation capacity within the aquifer,
and the sensitivity of the attenuation process to changes in
site conditions. The types of models that may be employed
as part of the site characterization process include simple
calculations, speciation models, reaction models, transport
models, and reactive transport models. Most modeling
undertaken in support of an application will be quantitative,
involving computer programs that require special skills to run
correctly. The contaminated natural system being modeled
is physically-, chemically-, and biologically-complex, and the
modeler must have a thorough knowledge of the processes
that affect the specific contaminants of concern. Site-specific
data collected to define the physical, chemical, and biological
characteristics of the aquifer are required to calibrate compo-
nents of the analytical models and test the validity of model
predictions. Deriving meaningful modeling results is likely to
require expenditure of significant amounts of time, and entail
considerable expense. This planning should occur early in
the site assessment process, so that the modeling can be
integrated with the evaluation of the site and the appropriate
data can be collected.
To obtain the best results at the least expense, it is important to
develop a valid modeling plan before beginning the modeling
itself. Developing such a plan will likely require the combined
talents of a group of specialists, including those familiar with
the site and those with expertise in applying quantitative
modeling of physical, chemical, and biological systems to
real-world problems. This section is devoted to giving general
perspective to the design and implementation of the modeling
strategy. In addition to the following discussion, the reader is
also referred to the document entitled "Documenting Ground-
Water Modeling at Sites Contaminated with Radioactive
Substances" (USEPA, 1996).
ID.1 Developing a Conceptual Model
Initially, the CSM is developed based on a general knowledge
of ground-water hydrogeology, ground-water geochemistry,
and known properties of the specific contaminant. With
acquisition of data that maps out the spatial and temporal
heterogeneity of the subsurface system, the CSM can be
updated. In general, there are more physical, chemical, and
biological processes operating in the subsurface of any given
site than can reasonably be accounted for in a modeling study.
The modeling effort begins with the careful identification of the
processes that play significant roles in contaminant migration
and attenuation at the site. In this way a conceptual model
emerges that will eventually be coded into the input streams
of the software packages that will produce the modeling re-
sults. If a correct and robust conceptual model is not derived,
the modeling results, no matter how detailed or expensive,
will contribute little to understanding the site, and will not be
supportive of the MNA application.
While it is important to begin modeling with a well-planned con-
ceptual model, the conceptual model may evolve as modeling
and collection of site-specific data proceeds. The processes of
observation and measurement and of modeling are, in practice,
closely interconnected. Initial observation and measurement
suggests a conceptual model, which supports development
of quantitative models. The results from application of these
quantitative models, in turn suggest additional important ob-
servations and measurements, which better constrain model
design and implementation. In this way, the conceptual model
is updated in an iterative fashion, as progressively more is
learned about the site. The most significant step in developing a
conceptual model of natural attenuation at the site is to identify
the transport and reaction mechanisms that significantly affect
the mobility of contaminants there. Once these mechanisms
have been identified, the logical components that will comprise
the conceptual model can be selected.
The evaluation of transport refers to analysis of the flow of
ground-water through the aquifer. The rate and direction of
ground-water flow will be governed by the physical characteris-
tics of the aquifer solids as well as the factors controlling inputs
of water into the aquifer. Spatial and temporal heterogeneity or
variability in these factors determines details of the mathemati-
cal construction of analytical models used to evaluate fluid and
contaminant migration through the aquifer. In characterizing
transport, it is important to ask questions such as:
• Does groundwater migrate through the bulk aquifer ma-
trix, through fractures or heterogeneities in the matrix, or
both?
• Does solute diffusion from areas of rapid flow to those
with stagnant conditions affect contaminant transport on
a scale finer than the envisioned numerical gridding, so
that a dual porosity model is required?
• Should the medium be considered homogeneous or het-
erogeneous on the scale envisioned for the nodal blocks
in the numerical gridding?
• Are medium properties best assigned deterministically, or
according to a stochastic algorithm?
• Is hydrodynamic dispersion described well in a Fickian
sense (i.e., in terms of dispersivity, according to Pick's
law), by differential advection through a numerical grid-
ding, or in both ways?
10
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• How can the model be calibrated to reflect as accurately
as possible transport rates through the subsurface?
• What additional data need to be collected to characterize
flow and calibrate the transport model? Such data might
include the distribution of hydraulic head, the evolution of
the contaminant plume through time, and the results of
tracer tests.
Evaluation of contaminant migration in ground water relies on
determination of the types of chemical reactions that control
contaminant degradation or immobilization. Thus, determina-
tion of specific reaction mechanisms that may be active within
a ground-water plume provides the basis for constructing
analytical models employed to evaluate performance of the
attenuation process and project contaminant transport into
the future. To characterize the reaction mechanisms driving
attenuation, it is necessary to ask questions such as:
• Does the contaminant adsorb to solid surfaces? If so, onto
what surfaces, and as what type of surface complex?
Does it desorb readily?
• Is the contaminant chemically oxidized or reduced? Is the
reaction catalyzed by mineral surfaces, or promoted by
microbial activity? If so, what is the catalyst or microbial
species?
• Does the contaminant precipitate as a solid phase? If it
does, what is the phase, and what is its solubility?
• Might complexation of the contaminant with chemical
constituents in solution affect its mobility?
A conceptual model can be thought of as a combination of
the logical components describing the various aspects of
transport and reaction at a site. For example, choice of how to
represent hydrodynamic dispersion, the equations to account
for sorption of contaminant species onto solid surfaces, rate
laws describing the kinetics of redox reactions, and equations
defining rates of microbial metabolism all contribute to the
conceptual model. Since a conceptual model is no more
than the sum of its components, and an analytical model is
simply the realization of a conceptual model, the final model-
ing results are no better than the components selected.
ID.2 Types of Models
There are several types of models that may prove useful for
characterizing attenuation processes at a site. In general,
in approaching a specific question, it is most expedient to
begin working with the simplest applicable model, adding
complexity to the study as necessary. It is wise to avoid the
temptation to begin by constructing the "ultimate" model, one
that accounts for all aspects of transport and reaction at a site.
Highly complex models are difficult to work with, expensive
to produce, and difficult to interpret. A more efficient strategy
is to begin with simple models of various aspects of the sys-
tem, combining these as necessary into progressively more
complex models, until reaching a satisfactory final result, one
that reproduces the salient aspects of the system's behavior
without introducing unnecessary complexity.
ID.2.1 Simple Calculations
Simple calculations performed by hand or via computer ap-
plications may provide an important component to the overall
modeling strategy. For purposes of this document, two
modeling approaches that fall under this category include
simplified calculation approaches to evaluate a range of
process outcomes and specific mathematical formulas used
to calculate input parameters needed for implementation of
more complex transport or reaction models. An example of
a simplified calculation approach would be the calculation of
the mass of contaminant and the mass of reactant within a
predefined volume of the aquifer for the purpose of assess-
ing if sufficient reactant mass is available for an identified
attenuation process. This type of calculation is simplified in
the sense that one may assume that the rate of the reaction
is unimportant. Thus, while this type of calculation provides
a general sense of the relative degree to which the aquifer
could support attenuation, it does not likely provide a suffi-
ciently accurate representation of the actual efficiency of the
attenuation process. However, the utility of this calculation
approach is to provide some perspective as to the relative im-
portance of investing resources to fully characterize reactant
mass or flux. Several examples of the second category of
this model type, specific mathematical formulas, are provided
at the following USEPA website - http://www.epa.gov/ath-
ens/learn2model/part-two/onsite/index.html. This website
provides on-line access to a suite of prepackaged tools (or
"calculators") for performing site assessment calculations.
Several examples relevant to site characterization advocated
within this document include:
• "Hydraulic Gradient Calculation" for assessing the
direction(s) of ground-water flow employing head
measurements in wells spaced horizontally across the
site;
• "Vertical Gradients" for assessing the potential for verti-
cal water transport within the aquifer based on head
measurements in closely-spaced, vertically nested
wells with identical screen lengths;
• "Vertical Gradients with Well Screen Effects" for assess-
ing the influence of variable screen lengths in vertically
nested wells on the calculated vertical gradient; and
• "Average Borehole Concentrations" to illustrate the po-
tential impact on contaminant concentrations measured
for samples collected from a single long-screened well
in an aquifer with a depth-varying concentration and a
depth-varying hydraulic conductivity field.
These simplified models support analysis of the adequacy
of the location and construction of ground-water wells,
which underpins the adequacy of the monitoring design
to provide samples and data reflective of the site-specific
conditions. They may also be used to provide reasonable
estimates for parameters needed as input to more complex
mass transport or reactive transport models. Since both
modeling approaches provide a means for preliminary as-
sessment of site data and potentially improving design of
the monitoring network, they play an important role in the
site characterization effort.
11
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ID.2.2 Mass Transport Models
Mass transport models seek to describe the flow of ground
water at a site, and the transport of chemical species within
the flow. Mass transport models are generally implemented
as transient simulations in one, two, or three dimensions.
Strictly speaking, a mass transport model considers the
migration of non-reacting species. In reality, many mass
transport codes can consider simple reaction scenarios, such
as partitioning of a species onto the solid surface accord-
ing to a constant partitioning factor. Mass transport models
can seldom be relied upon for describing natural attenuation,
because they lack sophisticated knowledge of chemical and
biological processes, but are nonetheless valuable in evaluating
a site's potential for MNA. The models are well developed
and straightforward to run; they are useful tools for simulating
the rate and pattern of groundwater flow at a site.
Mass transport modeling might be applied to figure the tran-
sit time of contaminants within the site, absent attenuating
processes. The models find use in applying the results of
tracer tests to calibrate the flow field. Some reactive transport
models (described below) accept externally determined flow
fields as input, so running a mass transport model may be a
required preliminary to a full reactive transport model.
ID.2.3 Speciation Models
Speciation models seek to describe the distribution of chemical
mass between solution, minerals, mineral surfaces, gases,
and biomass. Models of this class are useful because they
can predict the conditions under which contaminants might
be attenuated by sequestration, and those in which they are
likely to be mobile in the ground-water flow. For example, a
speciation model might demonstrate that a contaminant is
likely to adsorb to the surface of a component of the aquifer
solids over the pH range of interest. Or, the model might
show that the contaminant will tend to complex strongly with
dissolved chemical species, leaving it mobile and resistant
to attenuation.
Speciation models are implemented via the assumption that
the modeled system is in chemical equilibrium or, more com-
monly, partial chemical equilibrium. A model can be configured
to account for:
• Reactions among species in solution, including pro-
tonation-deprotonation, redox, and complexation reac-
tions.
• Adsorption reactions onto solid surfaces, possibly
including minerals and organic matter.
• Precipitation and dissolution reactions, to predict wheth-
er a mineral is saturated in solution, or undersaturated
or supersaturated.
• Gas solubility reactions, to account for the dissolution
of coexisting gases into solution, or the loss of gas
species from solution.
Where redox reactions play a critical role in the attenuation
reaction, it may be important to use a speciation model that
can account for redox disequilibrium. Microbial respiration, for
example, is driven by the transfer of electrons from donating
to accepting chemical constituents, including the inorganic con-
taminant. It may be critical, therefore, to characterize the redox
state of ground water at a site in an accurate and meaningful
manner to fully evaluate redox-driven reactions that influence
contaminant attenuation. Redox reactions in shallow ground
water rarely attain a state of equilibrium (e.g., Lindberg and
Runnells, 1984), which limits the utility of analytical models that
describe the distribution of chemical species in ground water
based on a single parameter such as dissolved oxygen (DO)
concentration or Eh (e.g., as measured using a DO or platinum
electrode, respectively). Geochemical models that describe
redox in terms of a single parameter may be limited in their
accuracy and/or flexibility in describing the redox characteristics
of the ground-water system. An alternative approach to the
model design would be to employ a flexible description of
redox in a state of chemical disequilibrium (e.g., as discussed
in Bethke, 1996, Chapters.). This type of modeling approach
allows the user to specify for each element the mass found
in the various possible redox states and reports the energy
(i.e., the Nernst Eh) associated with the half reaction for each
pairing of the element's oxidized and reduced states.
ID.2.4 Reaction Models
Reaction models are similar to speciation models in that they
consider the distribution of chemical mass, but have the addi-
tional ability of modeling the chemical evolution of the system.
Like speciation models, it is commonly necessary to use a
reaction model with a flexible description of redox disequilib-
rium, as well as suitable models to describe adsorption and
precipitation reactions. Where appropriate, the model should
be able to account for the kinetics of species sorption, redox
reactions, mineral precipitation and dissolution, or microbial
metabolism. Examples of the application of reaction models
in an MNA application include:
• Sequestration of contaminants onto a mineral surface as
the mineral forms, such as the complexation of heavy
metals in mine drainage onto ferrihydrite.
• Precipitation of contaminant-bearing minerals, according
to a kinetic rate law appropriate for the chemical condi-
tions at the site.
• Immobilization of a contaminant by oxidation or reduc-
tion, according to a kinetic rate law.
• Biotransformation of a contaminant by microbial life, us-
ing a rate equation for fermentation or cellular respiration
appropriate for conditions at the site.
ID.2.5 Reactive Transport Models
Reactive transport models, as the name suggests, are the
coupling of reaction models to transport models. Unlike a
reaction model, a reactive transport model predicts not only
the reactions that occur in the ground-water flow, but the dis-
tribution of those reactions across the site through time. A
reactive transport model of a site may have several advantages
over a simple reaction model, including:
• The ability to account for heterogeneity at the site, such as
an uneven distribution of a sorbing mineral, variation in
12
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pH conditions, or the differential development of microbial
populations.
• The ability to describe evolution of a contaminant plume
through space and time.
Reactive transport modeling is a relatively complex and time-
consuming undertaking, since it combines the data needs
and uncertainties inherent in modeling reaction as well as
transport, and because the calculation procedure may require
a significant amount of computing time. It may be the cap-
stone of the modeling effort, but is seldom the best tool for
initial scoping of the attenuation capabilities at a site. Such
modeling, on the other hand, may play an important role in the
site characterization effort, because it represents the integration
of all of the components of the conceptual model.
/D.3 Modeling and the Tiered Analysis
Approach
As described in Section 1C of this document, a tiered analy-
sis approach is recommended for organizing the collection
of site-specific data and providing a means for screening
out sites inappropriate for selection of MNA as part of the
ground-water remedy. Previously, possible applications of
models of varying complexity throughout the tiered analysis
process were provided in Table 1.1. The following discus-
sion provides additional context for evaluating the potential
role of model applications during the site characterization
process.
ID.3.1 Tier I - Demonstration of Contaminant
Removal from Ground Water
The application of models under Tier I pertains primar-
ily to initial characterization of hydrology and evaluating
whether measured ground-water characteristics may sup-
port immobilization processes. Assessment of hydrology
may include calculation of horizontal or vertical gradients
to assess the predominant direction(s) of ground-water
flow. This information could be used to guide installation
of monitoring points within the aquifer for collection of
ground-water and aquifer solids samples. Evaluation of
contaminant immobilization potential may involve use of
chemical data collected from ground-water and/or aquifer
solids samples as input into a speciation model to assess
the potential for contaminant precipitation or adsorption
onto aquifer solids. For example, speciation calculations
based on measurements of alkalinity and dissolved lead
within the ground-water plume may indicate saturation or
oversaturation with respect to precipitation of lead carbon-
ate. Conversely, measurements of ground-water chemistry
and extractable iron concentrations in aquifer solids could
serve as parameter inputs into a speciation model with the
capability of describing contaminant adsorption onto iron
oxides. It is recommended that these latter calculations
be used as secondary lines of evidence in support of site-
specific measurements that demonstrate active sorption of
the contaminant onto aquifer solids within the plume.
ID.3.2 Tier II - Determine Mechanism and Rate of
Attenuation
Modeling at this stage in the evaluation process should be
closely integrated with observational study. In studying the
mechanism of contaminant removal from ground water, care-
ful attention should be paid to assuring collection of sufficient
data to fully define the components of the conceptual model.
For example:
• If a precipitating phase is identified by x-ray diffraction,
spectroscopy, or electron microscopy, it will be necessary
to characterize the phase's solubility.
• If reaction with solid surfaces is identified as an important
attenuation process, it will be necessary to collect suf-
ficient data to properly parameterize an adsorption model
that describes the specific mechanism of adsorption, as
described in Section NIB.
• It may be necessary to establish a kinetic rate law describ-
ing precipitation of the contaminant into solid phases, or
its adsorption onto solid surfaces, where these reactions
may occur at different rates throughout the plume due to
the concentrations of aqueous or solid reactants.
In determining the rate of the attenuation process, modeling
may be used to describe chemical fluxes in the system and
rate of species uptake or production during chemical reaction.
Modeling might be specifically employed to estimate the time
frame required to sequester the contamination sufficiently to
meet cleanup objectives, where the attenuation reactions are
kinetically controlled.
/D.3.3 Tier III - Demonstrate Capacity and
Stability of Removal Mechanism
Model applications under Tier III would be directed toward
assessment of the capacity of the aquifer to attenuate the
mass of contaminant within the ground-water plume and the
long-term stability of an immobilized contaminant. Reaction
models and/or reactive transport models might be employed
to evaluate the extent of contaminant removal throughout
the plume. Use of these model types allows assessment of
rate-dependent reactions and/or the influence of decreases
in the flux of reactants due to changes in concentration or
ground-water flow that might occur over time. These same
models may be employed to evaluate ground-water conditions
that may remobilize contaminants sorbed to aquifer solids.
These evaluations may prove most useful for situations in
which laboratory testing may be less practical. For example,
model simulations may be employed to examine the stability
of the attenuated contaminant for hypothetical situations not
reflected in existing ambient ground water. For example, mod-
eling might be applied for a number of specific purposes:
• To test the chemical feasibility of specific remobilization
scenarios, such as infiltration of pristine groundwater, a
shift in oxidation state (perhaps due to waterlogging), or
a change in pH (due to soil acidification, for example).
13
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• To figure reactant fluxes required to remobilize the con-
taminant.
• To evaluate the possible effects of chelating agents, such
as organic acids, in the groundwater.
These model applications provide a means to project
system behavior under conditions that do not currently
exist, but could feasibly develop. They provide a source of
information that further reduces the uncertainty of reliance
on MNA as a permanent remedy.
ID.3.4 Tier IV - Long-Term Performance
Monitoring
Under Tier IV of the analysis process, modeling provides a
tool for designing a long-term monitoring plan, as well as a
contingency remedy for cases where unanticipated changes
in site conditions leads to failure of the MNA remedy. Modeling
tasks that might be performed at this stage include:
• Optimizing the location of monitoring wells for long-term
observation.
• Optimizing the frequency of sample collection events
based on knowledge of ground-water flow dynamics at
the site.
• Identifying critical chemical parameters to monitor based
on model simulations to examine the sensitivity of attenu-
ation process rate or capacity to changes in ground-water
composition.
• Identifying critical parameters to monitor based on model
simulations to evaluate conditions leading to contaminant
re-mobilization.
These model applications provide a means for designing the
monitoring program to best evaluate remedy performance
and provide site managers with a context for evaluating
possible decreases in the efficiency of the attenuation
process.
ID.4 Choosing Modeling Software
Once a modeling strategy has been developed and a con-
ceptual model defined, a computer software package (or
packages) will be needed to compute the modeling results.
A number of software packages exist for modeling physical,
chemical, and biological processes in natural systems. No
single package is best for all problems; one seeks the pack-
age or packages that best satisfies the objectives of the site
characterization process. Significantly, software packages
designed for analyzing problems of the MNA of organic con-
taminants (e.g., Bioplume III; USEPA, 1997) are generally not
suitable for studying the fate of inorganic contaminants. The
first step in selecting software involves identification of pack-
ages incorporating features needed to evaluate the conceptual
model. The selection process should amount to more than
compiling a checklist of features. It is important to determine
if the features work well for the situation in question.
It is critical to consider the efficiency of the software, not only
in computing time, but the time required to configure each
run and render the modeling results in a suitable graphical
form. One should, therefore, inspect carefully the documen-
tation from potentially suitable packages, and run test cases.
In evaluating a commercial package, insist on inspecting
the documentation before buying. Avoid licensing software
without being allowed a trial period, or a period during which
the software may be returned for a full refund.
ID.4.1 Public Domain vs. Commercial Software
Modeling software falls into two categories, public domain
and commercial. Public domain codes can generally be down-
loaded over the internet or purchased for a minimal charge;
some codes are obtained by personal request addressed to the
developer. A public domain code has a number of potential
advantages: there is little or no up-front cost; the source code
is in many cases available, allowing the modeler to correct
bugs and add features; and there may be a body of experi-
enced users available for consultation or troubleshooting at
minimal or no charge. A commercial code also has potential
advantages: it may be written by a group of professional pro-
grammers; there may be people assigned to support users,
offer training, and fix bugs; documentation may be superior;
there is more likely to be an intuitive user interface; the code
may be easier to use than public domain alternatives; and it
may offer superior graphics for rendering results. In general,
distributors of commercial codes hope they can convince
customers that the up-front costs of their product will be offset
in the long run by quality and savings, principally by improv-
ing the productivity of the people involved in the modeling
process, and by speeding project completion.
ID.4.2 Sources of Software
A considerable number of software packages that can be ap-
plied to the analysis of inorganic contaminant attenuation in
ground water are available in the public domain and from com-
mercial sources. Tables 1.2-1.4 list examples of various types
of commonly applied packages and their sources. Additional
packages may be found by searching the internet, and from
software retailers such as Rockware, Inc. (www.rockware.
com) and Scientific Software Group (www.scisoftware.com).
New software packages appear frequently, others fall into
disuse or are no longer supported and updated, and new
releases of the various packages add features and fix bugs.
As such, no attempt is made in this document to provide
exhaustive listings of software packages applicable to
MNA assessments, nor to judge the suitability or compile
the features of various packages. In evaluating software, the
reader will be well served by considering in light of his or her
own needs only the most recent available information. The
following discussion provides some issues to consider dur-
ing selection of a software package.
Issues to consider during selection of a mass transport
model and a representative list of commonly applied models
(Table 1.2):
• Whether the model operates in two or three dimensions,
or both.
• Whether the model can account for dispersion in the
manner chosen.
14
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• If the model accounts for saturated flow (flow below the
water table), unsaturated flow (above the water table),
or both.
• The deterministic or stochastic method or methods the
model can use to represent heterogeneity in the proper-
ties of the medium (hydraulic conductivity, dispersivity,
and so on) across the modeling domain.
Issues to consider during selection of a geochemical speciation
model and a representative list of commonly applied models
(Table 1.3):
• A flexible description of redox state. A disequilibrium
scheme in which each redox couple can be set to its
own redox potential is commonly required.
• The ability to account for sorption or surface complex-
ation in a manner appropriate for the site.
Issues to consider during selection of a reaction model, in
addition to those relevant for a speciation model, and a repre-
sentative list of commonly applied models (Table 1.3):
• An accounting for the kinetics of redox reactions,
whether occurring in the fluid phase, catalytically on
mineral surfaces, or promoted by enzymes.
• The ability to account for the kinetics of mineral pre-
cipitation and dissolution reactions invoked as an at-
tenuation mechanism, using appropriate rate laws.
• A model of microbial metabolism based on valid chemi-
cal principles. The metabolic model should treat the
Table 1.2 Example software packages for modeling groundwater flow and mass transport.
Software
FEFLOW
QMS
Modflow-2000
Visual Modflow
GroundWater Vistas
Source
Groundwater Modeling, Inc. www.ssg-int.com/
Environmental Modeling Systems, Inc. www.ems-i.com/GMS/gms.html
U. S. Geological Survey water.usgs.gov/nrp/gwsoftware/modflow.html
Waterloo Hydrogeology www.visual-modflow.com
www. grou nd water-vistas .com
Table 1.3 Example software packages for speciation in inorganic geochemical systems. Each of these packages
except Wateq4F also has at least some capability for modeling reaction processes.
Software
Chess
Eq3/6
Mineql+
MinteqA2
Phreeq-C
The Geochemist's
Workbench®
Visual Minteq
Wateq4F
Source
Ecole des Mines de Paris chess.ensmp.fr/
Lawrence Livermore National Laboratory
www.llnl.gov/IPandC/technology/software/softwaretitles/eq36.php
Environmental Research Software http://www.mineql.com/
U.S. EPA http://www.epa.gov/ceampubl/mmedia/minteq/
U.S. Geological Survey wwwbrr.cr.usgs.gov/projects/GWC coupled/phreeqc/index.html
University of Illinois www.geology.uiuc.edu/Hydrogeology
KTH (Sweden) www.lwr.kth.se/english/OurSoftware/Vminteq/index.htm
U.S. Geological Survey water.usgs.gov/software/wateq4f.html
15
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metabolism as a balanced chemical reaction, account-
ing for not only consumption of substrate species, but
generation of product species. The software should
also account for how the amount of energy available
in the environment affects metabolic rate, and for the
growth and decay of biomass.
Issues to consider during selection of a reactive transport
model, in addition to the points raised above about mass
transport, speciation, and reaction models, and a representa-
tive list of commonly applied models (Table 1.4):
• Whether the model can work in one, two, or three
dimensions.
• Compatibility of the model with the mass transport and
reaction models chosen. For example, can the reactive
transport model import a flow field predicted by the
mass transport model?
• Time to solution, since reactive transport modeling can
require considerable amounts of computing time.
ID.4.3 Th&rmodynamic Data
Most software packages are configured to accept any ex-
ternal database, provided that it is presented in the proper
format. A number of databases have been compiled for
various purposes, and many of these are available already
formatted to be read directly into one or more of the widely
distributed geochemical models. A list of various internet
sites from which thermodynamic data can be downloaded
in various formats is provided in Table 1.5. Additional data-
bases might be located by consulting web pages and the
latest documentation for the various geochemical modeling
packages, and by searching the internet. Since updates to
posted databases may be conducted infrequently, it may
be worthwhile to verify the database incorporates currently
accepted thermodynamic data based on a review of the
technical literature.
Table 1.4 Example software packages for modeling reactive transport in inorganic geochemical systems.
Software
Crunch
HYTEC
PHAST
Phreeq-C
The Geochemist's
Workbench®
Professional1
Source
Lawrence Livermore Laboratory www.csteefel.com/
Ecole des Mines de Paris www.cig.ensmp.fr/~vanderlee/hytec/index.html
U. S. Geological Survey
wwwbrr.cr.usgs.gov/projects/GWC coupled/phast/index.html
U.S. Geological Survey
wwwbrr.cr.usgs.gov/projects/GWC coupled/phreeqc/index.html
University of Illinois www.geology.uiuc.edu/Hydrogeology
1 The "Xt"package in previous releases.
Table 1.5 Example internet sources of thermodynamic data useful in constructing geochemical models.
Source
Ecole des Mines de Paris
Japan Nuclear Cycle Development Institute
Murdoch University (Australia)
National Institute of Standards and Technology
Nuclear Energy Agency (France)
University of Illinois
University of Illinois at Chicago
URL
ctdp.ensmp.fr/
migrationdb.inc.go.ip/
less, murdoch.edu. au/iess/i ess_home.htm
webbook.nist.gov/
www.nea.fr/html/dbtdb/
www.geology.uiuc.edu/Hydrogeology/hydro thermo.htm
tigger.uic.edu/~mansoori/TRL html
16
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/D.5 Accounting for Uncertainty
For a model constructed in support of an MNA application,
there are a number of sources of uncertainty, including:
• Error in chemical analyses. The accuracy and com-
pleteness of chemical analyses vary widely. Routine
chemical analyses performed by commercial laborato-
ries are in many cases of insufficient quality to support
geochemical and reactive transport modeling. Several
useful checks for internal consistency are available
in the American Water Works Association "Standard
Methods" volume (Clesceri et al., 1998), and computer
programs (e.g., Aq»QA, www.aqqa.com) are available
for performing these tests automatically. Geochemi-
cal modeling applications require complete chemical
analyses, including not only the contaminants of inter-
est, but the major ion chemistry, pH, and distribution
of metals among their mobile redox states.
• Error in determining hydrologic parameters. Measuring
representative values of hydrologic parameters such as
hydraulic conductivity and dispersivity can be difficult, be-
cause these values may change with the scale on which
they are observed. Laboratory measurements, therefore,
may give different results than well tests (e.g., slug and
bail tests, pumping tests), which may in turn differ from
values representative of the site as a whole. Measured
hydrologic parameters are important, but may need to
be calibrated to observations from the site, including
perhaps the rate of plume advance or the migration of a
tracer injected into the subsurface.
• Sample choice and dataset size. Significant error can
be introduced by sampling bias, although this bias
is not always obvious or even avoidable. Laboratory
measurements of hydrologic properties, for example,
are commonly made on samples that can be recovered
intact, even though the fractured or poorly consolidated
portions of the medium, left unsampled, control flow.
Fluid samples may be taken from monitoring wells
completed in highly conductive layers, where they can
be extracted rapidly, leaving unaccounted significant
quantities of residual contamination in slightly less con-
ductive layers. Finally, the number of samples available
or monitoring wells constructed is in some cases too
small to comprise a statistically significant dataset.
• Incompleteness and inaccuracy of the thermodynamic
database. To provide meaningful results, a geochemical
or reactive transport model has to include each of the
aqueous species, minerals, gases, and adsorbed species
important at the site, and the data for these species need
to be accurate. The thermodynamic databases available
for geochemical study vary widely in breadth and accu-
racy.
• Error in model components. Each of the components
of which the model is constructed is a potential source
of error. Components likely to contribute to error include
kinetic rate laws, surface complexation (sorption) models,
and descriptions of the effects of microbial metabolism.
• Conceptual errors. Perhaps most significantly, model
results can be affected by failure to conceptualize the
problem completely and accurately. If an important
process is not accounted for, or accounted for in an inac-
curate fashion, the modeling results will likely be rendered
useless.
The modeler accounts for uncertainty by experimenting with
the model to discover which sources of uncertainty affect
the results significantly. This uncertainty can subsequently
be reduced, for example, by making new measurements or
refining critical observations. Another source of uncertainty
is the limited possibility to obtain measured site-specific values
for some of the model parameters due to the complexity of
the geochemical model. It is recommended that the results
of uncertainty analysis be provided for the purpose of site
decisions. This information would include the sources and
potential ranges of all input data along with the origin of input
data (i.e., review of technical literature, model calibration, field
testing, or estimation).
/D.6 Mode/ Calibration and Verification
Developing a quantitative model of contaminant attenua-
tion in the subsurface may entail considerable uncertainty.
Parameters needed to constrain the model are seldom
known precisely, parameter inputs may not be available and
require estimation, and the conceptual model itself may need
refinement. Due to these uncertainties, it is necessary to
calibrate the model to observations, and to verify that the
model behaves in a manner that adequately describes the
natural system. The processes of calibration and verification
are closely related, since calibration brings the model into
alignment with observed data. A model that (1) utilizes to the
greatest extent possible parameter values specific to the site,
and (2) is calibrated to the observed evolution and distribution
of the contaminant plume, therefore, is most likely to be readily
verified. It is recommended that steps taken to calibrate the
model application be documented and provided for review in
order to build confidence in the use of this assessment tool.
Model verification requires that the model predict an inde-
pendent set of observations, i.e., a set separate from those
used for calibration. For example, a model that predicts the
attenuation of chromate by chemical reduction might be "fit"
on the basis of a plume or section thereof, and subsequently
used to predict the behavior of another plume at the same
site. The initial fitting would presumably involve arriving at
reasonably precise estimates of the most uncertain inputs - in
this case reduction rates, electron donor loads, and so on. If
the subsequent independent prediction accurately reflects field
observations, this result would lend credence to the model.
Here, "accurate reflection" of field predictions probably means
predicting correctly the speed at which the plume is retreating
and estimating the rate of overall contaminant mass reduction
to within a factor no greater than five. Predictions that do not
achieve this level of accuracy should prompt further refine-
ment of the model.
This discussion has been intended to point out that models
may serve as a useful tool that can be employed as part of
the evaluation process for selection of MNA as a remedy.
17
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However, the complexity of the modeling effort and the
potential level of uncertainty associated with model predic-
tions indicate that pursuit of more direct lines of evidence
is critical to the tiered analysis process. The acquisition of
these data will depend on establishing a network of moni-
toring locations throughout the aquifer. The site-specific
data collected from these monitoring locations provide
the means to identify the attenuation process and assess
the performance characteristics of the MNA remedy. As
with any technology used as part of a cleanup remedy,
continued assessment of remedy performance is critical
for ensuring attainment of cleanup goals. The following
discussion provides context for the eventual design of the
performance monitoring program leading to site closure for
situations in which MNA provides a viable component to
the ground-water remedy.
IE. Long-Term Performance Monitoring and
Site Closure
The performance of the MNA remedy must be monitored to
determine compliance with site-specific remedial objectives
identified in remedy decisions. This long-term monitoring
is often the largest expenditure incurred in the course of
cleanup and, for this reason alone, should be considered
at the earliest stages of remedial investigation. Because
the time horizons for successful implementation of an
MNA remedy are often expected to be long, it is critical
that particular attention is paid to long-term monitoring
plans. Detailed discussions of the performance monitoring
framework and monitoring plan development have recently
been published (USEPA, 2003). Although that discussion
focuses on attenuation of common organic contaminants,
the framework and many of the principles governing plan
development are also applicable to inorganic constituents.
However, there are conceptual differences with respect to
the outcome of the MNA remedy for inorganic contami-
nants. With the exception of situations where degradation
reactions transform harmful contaminants (e.g., nitrate
or perchlorate) into innocuous constituents, contaminant
mass is not reduced during MNA for inorganic contami-
nants. The MNA process results in relocation, dispersion,
and ultimately chemical conversion of the original source
zone. Therefore, the purposes of performance monitoring
are to demonstrate degradation to innocuous materials
and immobilization of contaminants. It is recommended
that site closure be considered only after degradation and
immobilization within the risk level specified in the remedy
decision are demonstrated and shown to have long term
stability.
Development of a performance monitoring plan is site
specific in nature. Monitoring objectives and quantifiable
performance criteria are developed to evaluate temporal
and spatial remedy performance with respect to the site-
specific remedial action objectives. Much of the monitoring
to demonstrate performance of the MNA remedy will fall
into three basic categories: 1) ambient monitoring to assess
background contaminant levels and the status of relevant
ambient geochemical indicators (e.g., EH, pH); 2) process
monitoring to assure the progress of chemical attenuation;
and 3) monitoring to detect plume expansion.
Within this framework, the OSWER Directive 9200.4-17P
(USEPA, 1999c) provides eight specific objectives to be met
by the performance monitoring program of an MNA remedy
(Table 1.6). The objectives usually will be met by imple-
menting a performance monitoring program that measures
contaminant concentrations, geochemical parameters, and
hydrologic parameters (e.g., hydraulic gradients). Much of
the monitoring will be focussed on ground water. However,
periodic monitoring of aquifer solids, through soil coring, will
be warranted in most situations. These data will be used
to evaluate the chemical behaviour of the contaminant in
the subsurface overtime, including:
• Changes in three-dimensional plume boundaries,
• Changes in the redox state that may indicate changes
in the rate and extent of natural attenuation,
• Reduction in the capacity of aquifer materials for con-
taminant immobilization, and
• Mobile contaminant mass and concentration reductions
indicative of progress toward contaminant removal
objectives.
Contaminant behavior can then be evaluated to judge the
effectiveness of the MNA remedy and the adequacy of the
monitoring program.
Table 1.6 Objectives for performance monitoring of MNA (USEPA, 1999c).
1)
2)
3)
4)
5)
6)
7)
8)
Demonstrate that natural attenuation is occurring according to expectations,
Detect changes in environmental conditions (e.g., hydrogeologic, geochemical, microbiological, or other
changes) that may reduce the efficacy of any of the natural attenuation processes,
Identify any potentially toxic and/or mobile transformation products,
Verify that the plume(s) is not expanding down gradient, laterally or vertically,
Verify no unacceptable impact to down gradient receptors,
Detect new releases of contaminants to the environment that could impact the effectiveness of the natural
attenuation remedy,
Demonstrate the efficacy of institutional controls that were put in place to protect potential receptors, and
Verify attainment of remediation objectives.
18
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IE.1 Duration and Monitoring Frequency
As stated in the OSWER Directive (USEPA, 1999c), per-
formance monitoring should continue until remediation
objectives have been achieved, and longer if necessary to
verify that the site no longer poses a threat to human health
or the environment. Typically, monitoring is continued for
a specified period after remediation objectives have been
achieved to ensure that concentration levels are stable and
remain below target levels. In order to demonstrate stability,
verification of the achievement of target levels under con-
ditions where the aquifer geochemistry has reestablished
a chemical steady state with respect to ambient ground-
water geochemistry will be needed. The magnitude of the
chemical gradient between the impacted and non-impacted
portions of the aquifer provides a reference point for evaluat-
ing establishment of steady-state conditions. A monitoring
strategy to verify the attainment of remedial objectives and
provide for termination of monitoring and site closure gen-
erally should be formulated during the development of the
performance monitoring plan and updated, as necessary,
prior to implementation.
Monitoring frequency should be specified in the perfor-
mance monitoring plan. In addition, the plan may specify
an approach and technical criteria that could be used to
increase or reduce the frequency as conditions change.
Such criteria would scale monitoring frequency to match
MNA performance and the level of understanding and
confidence in the conditions that control attenuation at a
given site. The most appropriate frequency for ground-
water sampling is site specific and depends on several
factors including:
• The rate at which contaminant concentrations may
change due to ground-water flow and natural attenua-
tion processes,
• The degree to which the causes of this variability are
known,
• The types of evaluations to be performed and the im-
portance of the type of data in question, and
• The location(s) of possible receptors relative to the
plume.
In addition, the most appropriate frequency may vary in
different areas of the site based on site-specific conditions
and the intended use of the data. Similar principles are
applied in determining the most appropriate frequency for
sampling of aquifer solids.
With respect to the initial frequency of ground-water sam-
pling under the performance monitoring program, quarterly
monitoring may often be an appropriate frequency to es-
tablish baseline conditions over a period of time sufficient
to observe seasonal trends, responses to recharge, and
to confirm attenuation rates for key contaminants. Quar-
terly monitoring for several years provides baseline data
to determine trends at new monitoring points and test key
hypotheses of the conceptual site model.
More frequent monitoring of ground-water elevations may
be warranted, particularly during the establishment of base-
line conditions, to improve the characterization of ground-
water flow patterns. In addition, more frequent monitoring
may be needed to observe changes in ground-water flow
patterns in response to other site activities, such as the
start or cessation of ground-water extraction in off-site water
supply wells, source control activities, and other significant
changes in the hydrologic system.
IE.2 Monitoring of Aquifer Solids
The aquifer material may serve as the reactive media to
which many inorganic contaminants become partitioned
and immobilized. Therefore, periodic re-assessment of the
capacity of aquifer materials for contaminant immobilization,
including immobilization of radioactive contaminants and
any harmful products of radioactive decay, often is a critical
step in performance monitoring. There are three aspects to
this solid-phase characterization to be addressed through
collection of field data and laboratory testing:
• Determination of the chemical process(es) resulting in
contaminant immobilization,
• Determination of the capacity of the un-reacted aquifer
material for contaminant immobilization, and
• Determination of the stability of the reacted aquifer
material with respect to contaminant release.
Characterization of aquifer material requires collection of
core material within the existing contaminant plume and
down gradient and side gradient to the plume. Charac-
terization within the existing plume is used to identify the
immobilization process(es) and capacity, while down gradi-
ent and side gradient characterization is used to re-assess
the potential and capacity for immobilization in the event
of plume expansion. In general, this characterization in-
volves identification of the aquifer mineralogy to determine
the abundance and spatial distribution of reactive solid
component(s) and the distribution of the contaminant among
the identified components.
The spatial extent and density of sampling points will be
dictated by the degree of heterogeneity of the aquifer
material both within and outside of the existing plume
boundary. The frequency of sampling will be dictated by
the rate of the immobilization process with respect to fluid
transport and the dynamics of fluid flow and chemistry. In
general, sampling frequency will be greater within the plume
boundary where immobilization is active. The frequency
of sampling outside of the plume boundary will be dictated
by the proximity of receptors and the time frame for reach-
ing remedial objectives relative to the rate of weathering
processes that may change the composition or mineralogy
of the aquifer material.
IE.3 Monitoring Types
The majority of the monitoring performed to determine the
effectiveness of the MNA remedy may be classified under
three general headings:
19
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• Monitoring of conditions outside of the plume boundar-
ies (ambient monitoring),
• Monitoring of natural attenuation processes (process
monitoring), and
• Monitoring to detect plume expansion and verify the
lack of impact to receptors (migration monitoring).
Other types of monitoring include periodic evaluations of
the effectiveness of any institutional controls specified in
the remedy decision documents and, ultimately, verification
of the attainment of all remedial objectives.
Ambient monitoring should be performed outside of the
boundaries (e.g., hydraulically up gradient, side gradient,
and down gradient) of the contaminant plume. The purpose
of this monitoring is to establish background conditions and
to provide an indication of the potential for additional plume
migration in situations where redox state and the capac-
ity of aquifer materials for contaminant immobilization are
dominant controls on migration. The extent and duration of
ambient monitoring will be influenced by the sensitivity of
aquifer chemistry to changes in recharge water quality and
processes that may change its composition.
Process monitoring is used to verify that attenuation is
occurring according to prediction. If process monitoring
indicates that attenuation is not occurring as expected, a
change in cleanup approach may be warranted. Process
monitoring is contaminant-specific and might include,
for example, measurement of ground-water redox state
or pH to assure the existence of conditions favorable for
natural attenuation via reduction-oxidation processes or
pH-dependent sorption as well as the monitoring of con-
taminants. Process monitoring parameters are discussed
in the contaminant-specific sections in Volumes 2 and 3 of
this document. Process monitoring should also take into
account any impacts of ongoing or prior active treatment on
subsequent ambient attenuation processes. For example,
such impacts may include gradual shifts in system redox as
water levels and/or electron donor/acceptor levels change
after, respectively, pump and treat or in situ bioremediation
have been halted.
Monitoring to detect plume expansion (migration monitoring)
and any impacts to receptors is another important aspect
of the performance monitoring program. This monitor-
ing objective may be met through multi-level monitoring
performed at or near the side gradient and down gradient
plume boundaries, beneath the plume, and near any other
compliance boundaries specified in remedy decision docu-
ments in conjunction with monitoring of possible receptor
locations (e.g., potable water wells or locations of ecological
receptors) to directly verify the lack of impacts. Monitoring
locations between the plume and compliance boundaries
or possible receptors should be close enough to the plume
that a contingency plan can be implemented before the con-
taminant can move past the point of compliance or impact
receptors. Identifying locations for monitoring wells de-
signed to detect migration ultimately relies on a site-specific
assessment of contaminant migration and fate. Additional
insight may be obtained from site-specific transport model
predictions, where model use is conducted iteratively with
the site characterization process so that model predictions
are both tested and influence future data collection.
IE.4 Monitoring Locations
At many sites, the performance monitoring program will
be three-dimensional in nature due in large measure to
the effects of site-specific hydrogeology on contaminant
migration. Typical target zones for monitoring a contaminant
plume (Figure 1.3) include:
• Original source areas - within and immediately down
gradient of source areas (Process Monitoring)
The monitoring objectives include the detection of any
further contaminant releases to ground water that may
occur and demonstration of reductions in contaminant con-
centrations in ground water over time. In situations where
the original source is contained, increased contamination
or new contaminants could be indicative of containment
system failure.
• Transmissive zones with highest contaminant concen-
trations or hydraulic conductivity (Process Monitor-
ing)
A change in conditions in these zones, such as an increase
in contaminant mass, change in redox state, increased
ground-water velocity, or exceedance of the aquifer capac-
ity for immobilization, may lead to relatively rapid plume
expansion.
• Distal or fringe portions within the plume (Process and
Migration Monitoring)
These are areas where reduction of contaminant concentra-
tions in ground water to levels required by remedial action
objectives may be attained most rapidly or where plume
expansion may be observed most readily.
• Outside the plume, including areas near plume bound-
aries and other compliance boundaries (Migration
Monitoring)
Multi-level monitoring points, reflecting vertical differences
in subsurface conditions, generally will be warranted at the
side gradient, down gradient, and vertical plume boundaries;
between these boundaries and possible receptors; and at
any other compliance boundaries specified in remedy deci-
sion documents. Monitoring of receptor locations should
also be included to directly verify that no impacts occur.
• Zones in which contaminant reductions in ground water
appear to be less than predicted (Process Monitor-
ing)
These are the areas where attaining cleanup standards
within time frames specified in the remedy decision docu-
ments may be impeded due to site conditions (e.g., higher
than anticipated concentrations of residual source materials,
redox conditions, or exceedance of the capacity for immobi-
lization). Such areas, if present, will be delineated through
evaluation of data obtained throughout the performance
20
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Site (Map View)
Upgradient
Transect
Lateral (Side Gradient)
Source
(T)Area
Recalcitrant
Zone
High
Concentration
Plume Core
Low
Concentration
Plume Fringe
Plume
Boundaries
Non-Hazardous
Degradation
Products & Other
Geochemical Indicators
Ground-Water Flow
Target Monitoring Zones
1. Source area
2. Contaminated zones of highest
concentrations and mobility
3. Plume fringes '
4. Plume boundaries ^
5. Recalcitrant zone determined
from historical trends
6. Upgradient and sidegradient
locations
Monitoring Well
Cluster
Monitoring well cluster
Piezometer
Transect of well clusters
Legend
Gravel, gravel-sand mixtures
Medium to coarse-grained sand
Fine-grained silty sand
• • Dissolved Plume
150
Figure 1.3 Example of a network design for performance monitoring, including target zones for monitoring effective-
ness with respect to specific remedial objectives.
monitoring period. These areas may require additional
characterization to determine if additional remedial actions
are necessary to reduce contaminant concentrations to
desired levels.
• Areas representative of uncontaminated settings (Ambi-
ent Monitoring)
Sampling locations for monitoring the redox state and im-
mobilization capacity of aquifer materials include points that
are adjacent to but outside the plume. Data from these
monitoring locations will often be needed to assess the
continuation of favorable conditions for attenuation. Since
assumptions concerning the redox state and attenuation
capacity affect interpretation of data from the plume, such
assumptions should be periodically evaluated like other
aspects of the conceptual site model. Therefore, multiple
monitoring points generally should be used to determine
the variability of these parameters outside the plume.
21
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• Areas supporting the monitoring of site hydrology
At some sites, monitoring of ground-water elevations
at locations additional to those used for the monitoring
of chemical parameters may be needed to determine if
changes in ground-water flow rates and directions are oc-
curring. Appropriate locations for placing piezometers will
often include positions that are up gradient, side gradient,
and down gradient of the contaminant plume, as well as
in zones above and below the plume and near surface
water bodies.
IE.5 Modification of the Performance
Monitoring Plan
The monitoring plan should be a dynamic document that is
modified as conditions change or the conceptual site model
is revised to reflect new information. Decisions regarding
remedy effectiveness and the adequacy of the monitoring
program will generally result in either:
• Continuation of the monitoring program without modi-
fication;
• Modification of the monitoring program;
• Implementation of a contingency or alternative remedy;
or
• Verification that remedial objectives have been met with
subsequent termination of the monitoring program.
Continuation of the program without modification would be
supported by contaminant concentrations behaving accord-
ing to remedial expectations while ground-water flow and
geochemical parameters remain within ranges indicative of
continued contaminant immobilization. Modification of the
program, including increases or decreases in monitoring
parameters, frequency, or locations, may be warranted to
reflect changing conditions or improved understanding of
natural attenuation processes at the site. In addition, modi-
fication generally would be warranted whenever remedy
modifications are implemented, such as implementation
of additional source removal or hydraulic control for plume
migration.
In situations where hydrologic and geochemical parameters
are stable and the contaminant concentrations in ground
water are decreasing as predicted, reductions in sampling
frequency (e.g., semi-annual, annual, or less frequent) will
often be warranted for process monitoring. For example,
five years of quarterly monitoring showing predictable
decreases in mobile contaminant concentrations might be
the basis for decreasing the frequency to a semi-annual
or annual basis at some sites. Ten years of semi-annual
or annual monitoring that shows predictable decreases in
mobile contaminant mass might likewise be the basis for
additional decreases in frequency, depending on site condi-
tions. Conversely, unexpected increases or lack of predicted
decreases in contaminant concentrations may trigger ad-
ditional characterization to determine the reasons for the
behavior, increased monitoring of pertinent parameters,
re-evaluation of the conceptual site model, and, potentially,
the implementation of a contingent or alternative remedy.
Changes in the frequency of monitoring to detect plume
expansion may also be warranted as process monitoring is
modified. However, the frequency of such monitoring should
not be decreased to the point where insufficient time would
be available for implementation of an effective contingency
plan in the event of MNA remedy failure.
Criteria for modifying the monitoring program, including the
type and amount of data needed to support the evaluation,
should be discussed and agreed to by stakeholders. Site-
specific criteria should be developed to define conditions
that indicate the appropriateness of increased or decreased
monitoring, additional characterization, re-evaluation of the
conceptual site model, implementation of a contingency
or alternative remedy, and termination of performance
monitoring.
Another reason for altering the monitoring program is the
development of more advanced monitoring technologies.
Because long-term monitoring costs are substantial, every
advantage of technological advances in monitoring ef-
ficiencies should be considered. This might best be done
by assessing monitoring technology every 3 to 5 years to
identify "off-the-shelf" monitoring approaches/equipment
that can improve accuracy and lower costs. National
technology verification programs are often a good source
of such information.
IE.6 Periodic Reassessment of Contaminant
Removal Technologies
In addition to the routine monitoring of MNA remedy per-
formance, it is recommended that periodic consideration
be given to any technological advances in the efficiencies
of source removal for inorganic contaminants. Implemen-
tation of more efficient technologies may result in reduc-
tions in the time frames for performance monitoring with
associated reductions in cost as well as improvements
in performance. Many sites may benefit from a Periodic
Remedial Technology Assessment (PRTA) conducted at
regular intervals (e.g., 5 years) throughout the performance
monitoring program. The PRTA should consist of a rigorous
literature search and engineering assessment of the field
implementation of new technologies. It should involve a
survey of cleanup efficiencies achieved by new technologies
at sites similar to the one under consideration. The survey
should rely on the results of national or state technology
verification programs (e.g., USEPA Environmental Tech-
nology Verification Program, www.epa.gov/etv/; Interstate
Technology & Regulatory Council, www.itrcweb.org). The
PRTA should either indicate the absence of more suitable
alternatives or suggest a faster path to site closure. The
criteria for technology selection should be clearly stated
during the development of the evaluation plan. The goal
of this review should be identification of technologies that
have a very high probability of achieving at least order-of-
magnitude reductions in contaminant mass and/or achieve-
ment of MCLs in ground water by means acceptable to
stakeholders. A reasonable metric should be successful
implementation of the technology as judged by impartial
bench marking criteria at several sites where site closure
has been achieved.
22
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and Applications, Oxford University Press, New York,
NY (1996).
Clesceri, L.S., A.E. Greenberg, and A.D. Eaton. Standard
Methods for the Examination of Water and Wastewa-
ter, 20th Edition, American Public Health Association,
Washington DC (1998).
USEPA. Environmental Characteristics of EPA, NRC, and
DOE Sites Contaminated with Radioactive Substances,
EPA 402-R-93-011, Office Radiation and Indoor Air,
Washington DC (1993).
USEPA. Common Chemicals Found at Superfund Sites,
EPA/OSWER No. 9203, Office of Solid Waste and
Emergency Response, Washington DC (1994).
USEPA. Contaminants and Remedial Options at Selected
Metal-Contaminated Sites, EPA/540/R-95/512, Wash-
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USEPA. Documenting Ground-Water Modeling at Sites
Contaminated with Radioactive Substances, EPA/540/
R-96/003, Office of Radiation and Indoor Air, Office of
Solid Waste and Emergency Response, Washington
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USEPA. Technical Protocol for Evaluating Natural At-
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ment, Washington DC (1998). (http://www.epa.gov/ada/
download/reports/protocol. pdf)
USEPA. Understanding Variation in Partition Coefficient, Kd,
Values: Volume I - Kd Model, Measurement Methods,
and Application of Chemical Reaction Codes, EPA
402-R-99-004A, Office of Radiation and Indoor Air,
Washington DC (1999a).
USEPA. Understanding Variation in Partition Coefficient,
Kd, Values: Volume II - Geochemisry and Available
Kd Values for Selected Inorganic Contaminants, EPA
402-R-99-004B, Office of Radiation and Indoor Air,
Washington DC (1999b).
USEPA. Use of Monitored Natural Attenuation at Superfund,
RCRA Corrective Action, and Underground Storage
Tank Sites, EPA/OSWER No. 9200.4-17P, Office of
Solid Waste and Emergency Response, Washington
DC(1999c).
USEPA. Evaluation of the Protocol for Natural Attenuation
of Chlorinated Solvents: Case Study at the Twin Cities
Army Ammunition Plant, EPA/600/R-01/025, Office of
Research and Development, Washington DC (2001).
(http://www.epa.gov/ada/download/reports/epa_600_
r01_025.pdf)
USEPA. Common Radionuclides Found at Superfund Sites,
EPA 540/R-00-004, Office of Radiation and Indoor Air,
Washington DC (2002a).
USEPA. Workshop on Monitoring Oxidation-Reduction
Processes for Ground-water Restoration, EPA/600/
R-02/002, Office of Research and Development,
Washington DC (2002b).
USEPA. Understanding Variation in Partition Coefficient,
Kd, Values: Volume III - Review of Geochemistry and
Available Kd Values for Americium, Arsenic, Curium,
Iodine, Neptunium, Radium, and Technetiurn, EPA
402-R-99-004C, Office of Radiation and Indoor Air,
Washington DC (2004).
23
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24
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Section II
Technical Basis for Natural Attenuation in Ground Water
Richard T. Wilkin, Steven Acree, Steve Mangion, Patrick V. Brady, Robert G. Ford,
Robert W. Puls, Paul M. Bertsch, Douglas B. Kent, Ann Azadpour-Keeley,
James E. Amonette, Craig Bethke
In determining whether MNA is applicable to a site, the
properties of the site and the properties of the contaminant
are analyzed in order to identify the specific process (or
processes) causing contaminant attenuation. Inorganic
contaminant transport in the subsurface will be governed by
the site-specific characteristics of ground-water flow and the
chemical interactions between the contaminant and aquifer
solids along the path (or paths) of fluid flow. The overall
extent of contaminant attenuation will be governed by the
velocity of ground-water flow relative to the rates of chemical
reactions that attenuate contaminant transport. The types
and rates of chemical reactions that result in contaminant
attenuation will be controlled by the availability of constitu-
ents within the aquifer that interact with the contaminant
in a manner that results in contaminant immobilization or
transformation. In simple terms, one can view the product
of the reaction between the contaminant and aquifer con-
stituents (or reactants) as the specific form of attenuation.
Examples of possible attenuation reactions are provided
in Figure 2.1 to illustrate this conceptual viewpoint and
the types of processes that may be active at a given site.
The reactants in the attenuation process may be present
in dissolved form or associated with aquifer solids (e.g.,
aquifer minerals or microbes). Thus, assessment for MNA
will necessitate collection of site-specific data that define
the processes controlling contaminant transport. In orderto
provide context for the types of data that may be required to
address this assessment objective, this section will provide
a review of the physical and biogeochemical processes that
govern contaminant transport in ground water.
Reactants
Attenuation
Contaminant + Aquifer Constituent(s) * Product(s)
Immobilization
PbCO
(GW) 3(GW)
3(S)
Cu(GW)+ goethite(S) * Cu-goethite(S)
Degradation
6NO,-..+ benzene,_+ 6H
3(GW)
,„...,- «, -,~,.,t microbe H
(GW) (GW) (GW/S)
3ISL + 6CCL + 6hUQ
(GW)
(GW)
(GW)
Figure 2.1 Conceptual view of attenuation as the interaction of the contaminant with aquifer constituents to form a
product resulting in attenuation. Subscript designations "(GW)"and "(S)"'indicate, respectively, whether
the reactant(s) and product(s) are in ground water or associated with immobile aquifer solids. For the
degradation reaction, N2 and CO2 are likely present as dissolved gases; "goethite" is the name of a com-
monly occurring iron oxyhydroxide mineral in aquifers.
25
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MA. Physical Transport Mechanisms
IIA.1 Basics of Ground-Water Flow and Solute
Movement
Ground water is rarely static and moves from areas of
ground-water recharge (i.e., high hydraulic head) to areas
of discharge (i.e., low hydraulic head). In a porous medium,
ground-water flow generally obeys Darcy's law with velocity
proportional to the hydraulic conductivity and hydraulic gra-
dient. In such settings, the average interstitial flow velocity,
or seepage velocity (Vs), is the rate at which water moves
through the pore spaces of the medium. Under natural
conditions, ground-water movement is relatively slow with
rates ranging from less than a foot per year to several feet
per day (USEPA, 1991). Seepage velocities for the various
aquifer materials at a site may often be estimated using:
Vs = Ki/ne
where,
K = the hydraulic conductivity of the medium
i = the magnitude of the hydraulic gradient
ne = the effective porosity or fraction of the medium
occupied by interconnected pore space.
In general, the seepage velocity at most sites is expected
to vary spatially due to heterogeneity in aquifer material
properties and temporally due to fluctuations in hydraulic
gradients.
The dominant processes that result in subsurface solute
movement in this dynamic environment are advection and
dispersion. Advection is the movement of a solute with the
bulk movement of ground water (Freeze and Cherry, 1979)
and generally is the primary solute transport mechanism
at sites with moderate to high ground-water flow rates.
Mechanical dispersion results in mixing of ground water
during advection reducing dissolved solute concentrations
within the plume and causing some solute molecules to
travel faster and some to travel slower than the average
ground-water velocity. A more detailed discussion of disper-
sion is provided by Gelhar (1993) and Gelhar et al. (1992).
The result at the macro scale is that a solute will spread
to occupy a larger portion of the flow field in the direction
of ground-water flow and in transverse (perpendicular)
directions than that due solely to advection. However,
limitations in the ability to obtain direct measurements of
dispersion relegate its determination primarily as a fitting
parameter during calibration of a ground-water flow model
for site-specific applications. At the field scale, stratification
(with associated differences in the hydraulic conductivity of
geologic materials) and fluctuations in hydraulic gradient
often result in much greater differences in the movement
of solutes relative to estimates based on the average
ground-water flow velocity (Figure 2.2). Thus, relative to
site-specific evaluation of contaminant transport, assess-
ment of the degree of variability in the distribution of hy-
draulic conductivity within the aquifer and time-dependent
changes in the magnitude and direction(s) of the hydraulic
gradient most likely plays a more critical role under the site
characterization effort.
Major Trend Due to
Velocity Distribution
Minor Spreading
by Dispersion
Figure 2.2 Cross-sectional view of differences in solute migration due to differences in hydraulic conductivity with
accompanying differences in ground-water velocity and the spreading of the solute front caused by dis-
persion (Keeley 1989).
26
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IIA.2 Colloidal Transport of Inorganic
Contaminants
The association of contaminants with suspended colloidal
material in ground water is a possible transport mechanism
and a complicating factor for accurate estimations of the
natural attenuation of contaminants in subsurface systems.
The mobile colloidal phase must be highly reactive, of suf-
ficient quantity, and stable for periods of time (i.e., remain
in suspension due to physical or chemical perturbations to
the aquifer) to enable the transport of a significant mass of
contaminants. Research to date indicates colloidal facili-
tated transport of contaminants in ground waters, surface
waters and in the unsaturated zone. Evidence suggests
that colloidal transport of contaminants may be significant
for some species under some hydrogeological conditions.
It is important therefore for sampling methods, transport
models, and site assessments to consider and be sensitive
to this transport mechanism.
Colloids are generally considered to be particles with
diameters less than 10 microns (Stumm and Morgan, 1981).
Based on theoretical considerations, Yao et al. (1971) and
O'Melia (1980) have estimated that the most mobile colloidal
particles in filtration studies in porous media may range from
0.1 to 1.0 urn. These include both organic and inorganic
materials. Recent estimates of colloidal concentrations
in ground water range as high as 63 mg L1 (Buddemeier
and Hunt, 1988), 60 mg L1 (Ryan and Gschwend, 1990),
and 20 mg L1 (Puls and Eychaner, 1990). In addition
to a high surface area per unit mass, colloidal particles
such as organic carbon particles, clay minerals and iron
oxides are also extremely reactive sorbents for inorganic
contaminants. If mobile in subsurface systems, these
colloids can effect the migration of contaminants for much
larger distances than many transport models would predict,
because sorbing contaminants spend a significant fraction
of time associated with mobile rather than immobile solids,
and because colloid transport can be primarily along large
diameter, fast flow paths.
Colloidal material may be released from the soil or geologic
matrix and transported large distances given favorable hy-
drological and geochemical conditions. Changes in solution
chemistry resulting from environmental pollution or changes
in ground water recharge chemistry can bring about
changes in the aqueous saturation state in the subsurface
leading to precipitation of new colloid-sized inorganic solids
that are entrained within flowing ground water, or can cause
the dissolution of matrix cementing agents, promoting the
release of colloid-sized particles. In addition, changes in
the concentrations of solutes that affect colloid surface
charge, such as pH or organic anions, can change the
stability of colloids. An excellent review of the mechanisms
of colloidal release, transport, and stability was published
by McCarthy and Zachara (1989). Reference to field and
laboratory studies provide examples of situations represen-
tative of sites with subsurface contamination where periods
of colloid mobilization may exist. For example, Gschwend
and Reynolds (1987) demonstrated that submicron ferrous
phosphate colloids were suspended and presumably mobile
in a sand and gravel aquifer. The colloids were formed from
sewage-derived phosphate and iron released from aquifer
solids due to reduction and dissolution of ferric iron from the
soil. Thompson et al. (2006) have also demonstrated that
microbially-driven redox cycling of iron bound to soils may
also lead to release of colloidal solids via indirect impacts
on water chemistry. In this instance, the mobilized colloidal
fraction was dominated by organic carbon solids.
However, field observations also point to the transience
of colloid mobilization in the subsurface. Nightingale and
Bianchi (1977) observed that ground-water turbidity may
increase for periods of time due to mobilization of colloidal
solids coincident with time-varying recharge events. These
observed turbidity increases were abated with time as the
aquifer returned to steady-state conditions. In addition,
Baumann et al. (2006) observed high colloid concentra-
tions in leachate from landfills, but colloid concentrations
decreased rapidly in ground water down gradient from the
landfill. Their observations suggested that the change of
hydrochemical conditions at the interface, from a reduc-
ing, high ionic strength environment inside of the disposal
sites to an oxidizing, low ionic strength environment in the
ground water (together with physical filtration effects for the
larger particles) was an effective chemical barrier for colloid
migration. Thus, it appears that while colloid mobilization
(along with associated contaminants) probably does oc-
cur, colloid migration is not likely to serve as a dominant
mechanism for contaminant migration encountered at
contaminated sites.
IIA.2.1 Implications for Natural Attenuation
Assessment
Many studies have documented that colloidal transport can
occur under some hydrogeological conditions. As noted
above, very few studies have shown that transport of con-
taminants via colloidal transport in the subsurface accounts
for the predominant mass of mobile contaminant at a site.
Where colloidal transport may be significant from a human
health standpoint is with highly toxic elements. From a site
characterization perspective, assessing the significance
of colloid-facilitated contaminant transport will depend on
the adequacy of well installation materials and construc-
tion approaches, as well as the approaches to sampling
ground water. Since well installation results in disturbance
of the subsurface solids, initial development to removed
fine-grained, disturbed aquifer materials from within and
adjacent to the well screen is a critical step prior to initiating
retrieval of ground-water samples. In addition, stabilization
criteria employed as indicators of the retrieval of represen-
tative ground-water samples may need to be adjusted if it
is important to rule out colloid transport as a contaminant
migration method. For this situation, more strict limits may
need to be placed on turbidity stabilization observations,
resulting in longer pumping times and/or significantly lower
pumping rates (e.g., Jensen and Christensen, 1999).
In addition, it is recommended that the chemical composi-
tion of the colloidal materials be identified at sites where
field data suggests colloid-facilitated contaminant transport.
27
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This step is warranted to rule out possible sources of sample
contamination or artifacts resulting from sample collection
or processing. For example, the presence of iron oxide col-
loids may be observed for reduced ground-water samples
containing elevated concentrations of ferrous iron [e.g., >5
ppm Fe(ll)] that are exposed to oxygen during collection
or handling. In this case, the likely source of colloidal iron
oxides is due to rapid oxidation and precipitation of ferrous
iron after the ground water has been pumped from the
subsurface. In general, the rigid and complex procedures
needed to insure sample quality for identification of ground-
water colloids may be difficult to implement within the overall
characterization effort (e.g., Dai et al., 2002).
Ultimately, the resources expended to this effort need to
be balanced with the effort to characterize site hydrology
as well as the subsurface processes that may control
contaminant degradation or sorption to aquifer solids. In
order to provide context to the types of data that may
be needed to identify specific sorption processes active
within the plume, the following sections provide detail on
the types of mechanisms that may result in contaminant
immobilization.
As previously outlined in Section I, the processes leading
to contaminant attenuation may include those that cause
reduction of contaminant mass (i.e., degradation and
radioactive decay) or cause immobilization of the con-
taminant via sorption to aquifer solids. For a majority of
the inorganic contaminants encountered at contaminated
sites, some form of immobilization will likely dominate the
attenuation process. The following discussion will provide
detail on the types of sorption processes that may result in
contaminant attenuation. This discussion will illustrate the
factors or parameters that are a component of the sorption
reaction in order to provide context to the types of measure-
ments needed to support identification of the site-specific
attenuation process and its performance characteristics
(discussed below in Section III of this volume). In addition,
the impact that microbial processes exert on the subsurface
geochemistry will be discussed in order to provide context
to factors that may dictate site-specific conditions within a
contaminant plume.
MB. Contaminant Sorption to Aquifer Solids
The primary, and in most instances, the only process in-
volved in the natural attenuation of inorganic contaminants
is through partitioning to the solid phase. The process of
contaminant transfer from the aqueous to the solid phase is
generally referred to as sorption and involves three primary
mechanisms (Sposito, 1986): adsorption, which is the ac-
cumulation of matter at the interface between the aqueous
phase and a solid adsorbent without the development of
a three-dimensional molecular arrangement; precipitation,
which is the growth of a solid phase exhibiting a molecular
unit that repeats itself in three dimensions; and absorption,
which is the diffusion of an aqueous or adsorbed chemical
species into a solid phase (Figure 2.3). From the stand-
point of monitored natural attenuation, the mechanisms
of most relevance will be dependent on contaminant and
site specific characteristics, such as the surface reactivity,
solubility, and redox sensitivity of the contaminant, as well
as the type and abundance of reactive mineral phases and
the ground-water chemistry.
More detailed discussion is provided below on the various
sorption processes introduced in the preceding paragraph.
As a point of reference, representative examples of sorp-
tion processes for specific contaminants are provided in
Figure 2.4.
Diffuse Ion
contaminant
coordinated H2O
i
™ 1*
. . •
Aquifer
Mineral
o
(d)
Figure 2.3 Representation of an aquifer mineral surface with (a) an outer-sphere surface complex; (b) an inner-
sphere surface complex; (c) a multinuclear surface complex or a surface precipitate; and (d) absorption,
or solid state diffusion and substitution of the sorbate in the mineral structure (after Sposito, 1984).
28
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IIB.1 Adsorption
Adsorption processes are typically categorized by the rela-
tive binding strength of interaction between the adsorbate
(species in solution) and the adsorbent (aquifer solid).
There is a range of binding strength for contaminant ad-
sorption that depends on characteristics of the adsorbate,
sorbent, and ground-water chemistry. However, discussions
of binding strength are generally couched in terms of "weak"
or "strong" adsorption processes, albeit a common conven-
tion in chemistry would categorize both the covalent and
electrostatic interactions involved in adsorption as 'strong'
intermolecular forces (Israelachvili ,1994). One microscopic
distinction borrowed from the characterization of soluble
ion pairs that is commonly used to delineate weak and
strong adsorption involves the solvation properties of the
adsorbate (Westall, 1986; Stumm, 1992). If solvating water
molecules are interposed between the cation or anion and
the surface, the adsorption complex is referred to as outer
sphere and is considered to be weak. Conversely, if upon
adsorption the adsorbate loses waters of hydration such
that there are no water molecules interposed between the
cation or anion and the surface, the adsorption complex is
referred to as inner sphere and is considered to be strong
(Sposito, 1984). The propensity of a cation or anion to form
either an inner-sphere or outer-sphere surface complex is
a function of the adsorbate, the surface functional groups
of the adsorbent, and aqueous phase chemistry (e.g., pH
and ionic strength).
IIB.1.1 Reactive Mineral Phases Involved in
Adsorption
Important adsorbent phases commonly found in the
environment include phyllosilicate minerals, metal
oxyhydroxide phases, sulfide phases, and natural organic
matter (Dixon and Schulze, 2002). Many phyllosilicate
minerals possess a permanent negative charge as a result
of the substitution of lower valence cations, i.e., Mg(ll),
Fe(ll), Li(l) for Al(lll) in the octahedral layer and/or Al(lll)
for Si(IV) in the tetrahedral layer (referred to as isomorphic
substitution). There are two main classes of phyllosilicate
minerals based on layer structure (Figure 2.5). The 1:1
mineral layer type is comprised of one Si tetrahedral layer
and one Al octahedral layer, which in soils and aquifers
is commonly represented by the mineral kaolinite having
the general formula [Si4]AI4O10(OH)8»nH2O. Kaolinite and
related minerals generally have insignificant degrees of
cation substitution within their octahedral and tetrahedral
layers, and, thus generally posses a very low permanent
negative charge. The 2:1 mineral type is comprised of one
Al octahedral layer interposed between two Si tetrahedral
layers comparable to the mica structures (Figure 2.5). The
2:1 layer class is represented by a variety of minerals,
which are classified based on the location (tetrahedral
vs. octahedral layer) and relative amount of isomorphic
substitution. The three major mineral classes within the 2:1
layer type are illite (Mx[Si68AI12](AI3Fe025Mg075)O20(OH)4),
vermiculite (Mx[Si7AI](AI3Fe05Mg05)O20(OH)4), and smectite
(Mx[Si8]AI32Fe02Mg06O20(OH)4), which display different levels
Mobile
Contaminant
Reactant
Product
Immobile
Precipitation
Aqueous
I
Solid
Coprecipitation
Aqueous
Solid
Adsorption
Figure 2.4 Examples of contaminant-specific sorption processes that may lead to attenuation of the ground-water
plume. Color coding is employed to distinguish the contaminant (red), aqueous or solid phase reactants
(blue), and the product (yellow) of the reaction leading to contaminant attenuation. Absorption is illustrat-
ed as a possible sequential process that follows the adsorption of a contaminant onto a mineral compo-
nent within aquifer solids.
29
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of cation substitution in their tetrahedral and octahedral
layers. The permanent negative charge imparted to 2:1 clay
minerals by isomorphic substitution is typically balanced
through exchange reactions involving major cations in
ground water (e.g., Na+, K+, Ca2+, or Mg2+; represented by
"Mx" in the formulas listed above).
Contaminant sorption to phyllosilicates may occur via ion
exchange or surface complexation with surface functional
groups (see examples labeled "surface complex" and "ion
exchange" in Figure 2.5). Due to differences in the levels
of isomorphic substitution for the 1:1 and 2:1 clay mineral
classes, ion exchange is usually only significant for 2:1 phyl-
losilicates. In addition to siloxane oxygen atoms along the
basal plane, phyllosilicates possess two types of terminal
ionizable OH groups, aluminol and silanol, protruding from
the edge surface. These edge OH groups can form both
inner- and outer-sphere complexes with metal cations and
oxyanions depending on the pH of the bathing solution and
on the specific characteristics of the cation or oxyanion
(represented as "surface complex" in Figure 2.5).
The most important surface reactive phases for both cat-
ionic and anionic contaminants in many soil and subsurface
systems are the metal oxyhydroxide phases. These phases
are characterized by hexagonal or cubic close-packed O or
OH anions with Fe2+3+, AI3+, and/or Mn3+4+ occupying octa-
hedral sites. These oxides are present as discrete phases
and as complex mineral assemblages, being co-associated
with phyllosilicates and primary minerals as coatings or
with humic macromolecules. In soils and sediments the
crystallinity of these phases typically varies from poorly
ordered to well crystalline forms and grain size from the
nanometer to micrometer scale. Among the most common
Fe-oxyhydroxide phases found in soils and sediments are
the poorly ordered phase ferrihydrite (Fe2O3 nH2O), and the
moderate to well crystalline phases, goethite (a-FeOOH),
and hematite (a-Fe2O3). The most common Al oxyhydroxide
phase found in soils and sediments is gibbsite (y-AI(OH)3).
Additionally, poorly ordered aluminosilicates can be im-
portant reactive phases in certain soils and these include
the very poorly ordered allophanes (Si/AI ratios 1:2 to 1:1)
and the paracyrstalline phase, imogolite (SiO2 AI2O32H2O).
While Mn oxyhydroxides are less prevalent than Fe- and
Al-oxyhydroxides in soils and sediments they are very im-
portant phases in terms of surface mediated redox reactions
and because of their propensity for high metal sorption.
The mineralogy of Mn is complicated by the range in Mn-
O bond lengths resulting from extensive substitution of of
Mn2+ and Mn3+ for Mn(IV). Thus, there exists a continuous
series of stable and metastable compositions from MnO to
MnO2 forming a large variety of minerals. Among the more
common Mn-oxyhydroxides are pyrolusite (|3-MnO2), the
hollandite-cryptomelane family (a-MnO2), todorokite, and
birnessite (a-MnO2).
1:1 Mineral Type
(Kaolinite)
2:1 Mineral Type
(Smectite)
Sorbed
Contaminant
Clay Mineral
O Oxygen
® Hydroxyl
• Aluminum (Iron, Magnesium)
• o Silicon (Aluminum)
^ Cesium
Ozinc
Figure 2.5 Diagrammatic sketch of the structure of 1:1 and 2:1 phyllosilicate minerals. Also shown are hypothetical
sorption reactions for zinc and cesium (ion exchange represented as 'Mx'in structural formulas for 2:1
phyllosilicates shown in text).
30
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IIB.1.2 Surface Functional Groups on Aquifer
Solids and the Impact on Surface Charge
The reactive surface functional group for all of the metal
oxyhydroxide phases is the inorganic OH moiety exposed
on the outer periphery of these minerals. The reactivity of
a specific metal oxyhydroxide is dependent on the surface
area (SA), surface-site density (Ns), the degree of coordina-
tion of the OH group to the bulk structure, and the point of
zero charge (PZC).The charge on the mineral surface may
impose either attractive or repulsive contributions to the
overall adsorption reaction, depending on the type of charge
possessed by the adsorbate. The properties of the sorbent
that impact adsorption are controlled by both the grain size
and specific structure of the oxyhydroxide phase.
The surface charge of oxyhydroxide minerals and edge
sites on phyllosilicates is derived from the protonation and
ionization of exposed surface hydroxyl groups, represented
by = SOHn"'1, where S represents the structural metal cation
(e.g., Fe, Al, Mn) over a stoichiometric range from n= 0, 1,
or 2. Thus, as a function of pH, the surface functional groups
can be generally described with the following idealized no-
menclature: = SOH2+, = SOH°, and = SO. The exact charge
associated with the various surface functional groups is
difficult to measure, so the main purpose of employing this
nomenclature is to illustrate that surface charge varies as
a function of ground-water chemistry. The gradual change
in surface charging with pH for some common minerals is
illustrated in Figure 2.6 and a discussion of surface site
charging is provided below.
Natural organic matter comprised of complex polymers
called humic substances, represents another very impor-
tant reactive phase in aquifer solids. A variety of functional
groups are present in humic substances, and, like OH
functional groups of the inorganic metal oxyhydroxides,
these also are characterized by pH dependent charging
mechanisms. The primary functional groups associated with
humic substances in terms of surface charge are carboxyl
and phenolic groups, however the less abundant amino,
imidazole, and sulfhydryl groups may play an important
role in the sorption of certain contaminant metals when
present at trace levels (Table 2.1).
Based on the previous discussion, it is apparent that the
charge on aquifer solids can be grouped into two classes
associated with the mechanisms that give rise to electrical
charge associated with mineral surfaces or with organic
functional groups. These two classes are commonly re-
ferred to as permanent (or constant) charge and variable
charge.
• Constant charge - Constant charge is the predominant
charge in phyllosilicate clays. Because, for the most
part, these isomorphic substitutions occur during
mineral formation, this charge deficit is fixed in the lat-
tice structure and is hence unaffected by changes in
electrolyte concentration or pH of the soil solution.
• Variable (pH dependent) charge - Variable charge is
the predominant charge for oxyhydroxide minerals
such as hematite, goethite, and gibbsite, as well as
20 r
12
12
16
20 I-
Ferrihydrite
Goethite
Figure 2.6 Surface charge of some hydroxides from
pH 2 to 10 measured in different electrolyte
solutions shown in parentheses; positive and
negative surface charge shown above and
below the x-axis, respectively. Ferrihydrite
[Fe(OH)3nH2O] (0.001 M NaNOJ from Hsi
and Langmuir (1985); gibbsite [AlfOHjJ and
silica gel [SiO2nH2O] (1.0 M CsCI) based on
Greenland and Mott (1978); goethite [OL-
FeOOH] (0.005 M CsCI) based on Green-
land and Mott (1978) (see also Hsi, 1981);
birnessite (o-MnOJ (0.001 M A/aA/Cy based
on Caffs and Langmuir (1986).
for humic substances. The metal ions in the vicinity of
the surface of metal oxyhydroxide minerals are coor-
dinatatively unsaturated, i.e. they are lewis acids, and
coordinate with water molecules, which subsequently
dissociate a proton leading to a surface layer of metal
hydroxide functional groups. This process also occurs
at the edges of phyllosilicate clays giving rise to SiOH
(silanol) and AIOH (aluminol) functional groups. These
surface hydroxyl groups can become positively or
negatively charged by binding or dissociating a proton
(i.e., protonation-deprotonation reactions):
S—OH + H+
S—OH
S—OH2
S—O-+H+
Thus, the prevalent surface charge in aquifer solids will
be dependent on the pH of ground water and the types
and concentrations of ions that balance the permanent
charge association with phyllosilicates. The extent to which
31
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Table 2.1 Important functional groups in humic sub-
stances that impact surface charging behav-
ior and contaminant binding.
Functional Group
Amino
Carboxyl
Carbonyl
Alcoholic hydroxyl
Phenolic hydroxyl
Imidazole
Sulfhydryl
Structural Formula
-NH3
o
— C — OH
0
— C — NH
-OH
CH
(( )VNH
-SH
protonation or deprotonation occurs is also a function of the
metal ion and the local binding environment of the metal
hydroxide surface group. The highly electropositive Si4+ in
silanols precludes the protonation of the surface hydroxyl
and this functional group can only dissociate a proton
under pH conditions generally encountered in ground water.
Aluminols, on the other hand, can be either positively or
negatively charged. Various types of hydroxyls of differing
reactivity have been identified spectroscopically at the
surface of metal oxides.
The charge of aquifer minerals is always electrically bal-
anced by interactions with ions of an opposite charge
(counter-ions). We can define two broad classes of weak
and strong interactions (outer and inner-sphere) that act to
neutralize the charge developed at soil mineral surfaces.
IIB.1.3 Weak and Strong Adsorption Regimes
Weak adsorption regime - Within the weak adsorption re-
gime, simple ion exchange is the most common mechanism
and involves the electrostatic attraction of an ionic species
by a negative or positive charge emanating from a mineral
surface or from functional groups associated with humic
substances (Sposito, 1981). Long before the structures of
reactive soil minerals were determined, it was observed
that, under certain circumstances, there was a reversible
and stoichiometric (based on charge) replacement of major
cations in soils equilibrated with concentrated neutral salt
solutions according to the reaction:
Na2-Xs
Ca2
Ca-X +2Na+
Soil and sediment materials are typically characterized by
their cation exchange capacity (CEC), which is defined as
the quantity of cations reversibly adsorbed per unit weight
of mineral and typically expressed as cmol kg'1. The cation
exchange capacity of 2:1 phyllosilicate clays tends to be
constant over a wide pH range, since ionizable edge groups
are relatively minor on a surface area basis compared to
the permanent charge associated with planar sites. For
2:1 phyllosilicates, cations hydrated to differing degrees
are located in the interlayer space and can be displaced
by other competing cations through ion exchange reactions
(see Figure 2.5). In principle, cation exchange reactions
involve both inner and outer sphere complexation with pla-
nar sites, although except for the special case discussed
below for large weakly hydrated monovalent cations, such
as K+ and Cs+, both are readily reversible. Both inner and
outer sphere complexes can also form with O functional
groups associated with organic macromolecules and O and
OH atoms associated with metal oxyhydroxides, but only
the outer sphere complexes are considered weak adsorp-
tion. The major difference between phyllosilicates having
substantial isomorphic substitution and metal oxyhydroxides
and humic substances, is that the CEC is highly pH depen-
dent, increasing with increasing pH. Since reactive mineral
phases in soils and sediments are a composed of complex
assemblages of phyllosilicates, oxyhydroxides, and humic
substances, CEC is always a pH dependent property.
Strong adsorption regime - As discussed above, simple ion
exchange is the predominant adsorption mechanism for
phyllosilicate clays. A major exception to this is the inner-
sphere sorption of larger unhydrated cations, such as K+
and Cs+ to oxygen atoms of two opposing siloxane ditrigonal
cavities of collapsed layers of weathered micaceous miner-
als, such as illite, which can be classified as an 'irreversible'
adsorption or as an absorption process.
At this point it is important to discuss the concept of the
reversibility of adsorption. From the perspective of chemical
thermodynamics, the definition of a 'reversible' process is
one where the initial state of the system can be restored
with no observable effects in the system and its surround-
ings (Holman, 1980). The use of the term 'irreversible'
from the standpoint of adsorption mechanisms is relative
and does not strictly adhere to the thermodynamic (or
chemical) definition in all cases. The fixation of Cs+ in illitic
minerals is conceptually thought to proceed via an initial
ion exchange reaction followed by an interlayer collapse
(fast) or through the slower migration into interlayer sites
in collapsed layers (absorption). Once Cs+ is fixed within
the interlayer of the clay mineral, its release is not readily
reversible via displacement with competing solutes, i.e.,
through ion exchange mechanisms. Thus, the release of
fixed Cs+ is subsequently controlled by a process such as
mineral dissolution. In this sense, the original fixation pro-
cess is irreversible, since contaminant release would result
from a mechanism other than the reversal of the original
adsorption mechanism.
The formation of a chemical bond between an adsorbate
and a functional group on the adsorbent also falls within
the category of a strong adsorption regime. In general,
breaking chemical bonds requires more energy than over-
coming electrostatic interactions. Metal adsorption to OH
functional groups of oxyhydroxide phases through surface
complexation can be illustrated by the following surface
reaction (Stumm, 1992; McBride, 2000):
=Fe-OH'1
M(H.O)6
=Fe-O-M(H2O)5<"-3/2>+ + H3O+
32
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Thus, specific adsorption of cations increases the positive
charge at the mineral surface when n > 1, which is gener-
ally the case for transition elements, and results in the net
release of H+ ions to the soil solution. Adsorption of anions
from solution occurs by ligand exchange of a OH or H2O at
the surface functional group according the following reaction
(solution anion represented by A""):
=Fe-OH
+OH-
Anion adsorption is favored by low pH, which leads to
protonation of surface functional groups and makes them
better leaving groups in the ligand exchange reaction.
IIB.2 Precipitation
Mineral-water reactions occur as ground water moves
through porous media. These reactions may result in the
removal of aquifer components due to mineral dissolution
or result in the buildup to oversaturation and consequent
precipitation of secondary minerals. As an outcome of
mineral-water reactions along a flow path, fluid composi-
tions and the mineralogical makeup of the solid phase will
continuously evolve towards a stable or equilibrium state.
Mineral precipitation processes in aquifer systems are an
important group of immobilization mechanisms for inorganic
contaminants in ground water.
Full treatment of precipitation processes, including cover-
age of relevant thermodynamic and kinetic concepts, is
outside the scope of this document. The reader is referred
to numerous standard textbooks in geochemistry, soil sci-
ence, and aquatic chemistry (e.g., Lindsay, 1979; Stumm,
and Morgan, 1981; Drever, 1982; Sposito, 1989; Stumm
1992; Morel and Hering, 1993; McBride, 1994; Sparks,
1995; Langmuir, 1997; Lasaga, 1999). The purpose of this
section is to introduce key concepts and issues regarding
the potential impact precipitation reactions may exert on
contaminate attenuation. In general, mineral precipitation
in relation to contaminant immobilization can be discussed
in the context of four widely studied processes:
• Precipitation from solution: Nucleation and growth of
a solid phase exhibiting a molecular unit that repeats
itself in three dimensions. Homogeneous nucleation
occurs from bulk solution and heterogeneous nucleation
occurs on the surfaces of organic or mineral particles.
Heterogeneous nucleation is thought to be more impor-
tant in natural systems that are rich in reactive inorganic
and biological surfaces. Precipitation may result in the
formation of sparingly soluble hydroxides, carbonates,
and, in anoxic systems, sulfides. Many precipitation
reactions have a strong dependence on pH.
• Coprecipitation: Incorporation of an element as a trace
or minor constituent within a precipitating phase. In this
case, the contaminant substitutes for a more concen-
trated component in the crystal lattice (isomorphous
substitution). This process is distinct from adsorption
due to incorporation of the contaminant within the bulk
structure of the major mineral phase. Examples of
coprecipitation include Cr(lll) in hydrous ferric oxide,
Cd(ll) in calcium carbonate, and As(lll) in iron sulfide.
• Surface precipitation: A precipitation process interme-
diate between surface complexation and precipitation
from bulk solution. Surface precipitation represents the
continuous growth of particles formed via heteroge-
neous nucleation. Macroscopic studies of adsorption
of some solutes, particularly di-valent and tri-valent
cations, suggest that precipitation occurs at surfaces
under conditions where the solid is apparently under-
saturated based on solution concentrations (Dzombak
and Morel, 1990).
• Mineral transformation: Adsorbed contaminants can
become incorporated into minerals that form as a result
of recrystallization or mineral transformation processes
in soils and sediments. Transformation reactions may
be accelerated or retarded by the contaminant, and in
some cases mineral transformation may result in the
exclusion of the impurity contaminant from the solid
phase. Examples include incorporation of anions, such
as As(V), into hydrous ferric oxide and transformation
to Fe oxyhydroxides (e.g., Ford, 2002), coprecipitation
of metals with iron monosulfide and transformation to
iron disulfide (e.g., Lowers et al., 2007), and layered
double hydroxides (typically with Al) as intermediates
between adsorbed/surface precipitated metal ions like
Ni and Zn, and metal-ion-containing aluminosilicates
(e.g., Ford, 2007).
The relative importance of these processes will be deter-
mined by contaminant characteristics as well as site-specific
characteristics of the plume ground-water chemistry and
aquifer solids. These individual processes are discussed
in more detail below.
IIB.2.1 Precipitation from Solution
Solution precipitation or crystallization can be divided
into two main processes: nucleation and crystal growth.
Nucleation occurs prior to growth of a mineral crystal. Both
nucleation and growth processes require a system to be
oversaturated in the new phase. The probability of nucle-
ation occurring increases exponentially with the degree of
oversaturation. Nucleation of a new phase is often facili-
tated in the presence of a surface (heterogeneous nucle-
ation) compared to bulk solution (homogeneous nucleation).
Because nucleation and growth are processes that compete
for dissolved solutes, at high degrees of oversaturation the
rate of nucleation may be so fast that all excess solute is
partitioned into crystal nuclei. In contrast, lower levels of
oversaturation can result in the growth of existing crystals
without nucleation. Well-formed or euhedral crystals typi-
cally develop slowly via growth from solution at low degrees
of oversaturation. During crystal growth various chemical
reactions can occur at the surface of the growing mineral,
such as adsorption, ion exchange, diffusion, and forma-
tion of surface precipitates. In general, the rate of crystal
growth is controlled either by transport of solutes to the
growing surface (i.e., transport controlled), by reactions at
the surface (i.e., surface controlled), or a combination of
these factors.
33
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For the most abundant cations present in aquifers and
soils, such as Al, Si, Fe, Mn, Ca, and Mg, precipitation of
mineral forms is common and will in many cases control
concentrations observed in solution. At contaminated sites,
concentrations of ground-water contaminants are typically
several orders of magnitude below the concentrations of the
dominant solutes in water. At low concentrations, sorption,
surface precipitation, or formation of a dilute solid solution
(coprecipitation) may be the more probable removal pro-
cesses for contaminant species (McBride, 1994). However,
precipitation of iron-bearing or aluminum-bearing minerals,
for example, can have an important affect on the transport
and fate of metal and metalloid contaminants. Major mineral
precipitate classes encountered in aquifers and soils are
listed in Table 2.2.
The tendency for a system to support a specific precipitation
or dissolution reaction can be evaluated through compari-
son of the equilibrium solubility constant for a given solid
phase to the ion activity product calculated using ground-
water chemical data. The relative magnitude of the values
of the equilibrium solubility product and the calculated ion
activity product provides a measure of the saturation state of
ground water relative to mineral precipitation or dissolution.
A conventional method for expressing the ground-water
saturation state is by calculation of the saturation index,
SI, which is given by (Stumm and Morgan, 1981):
SI = AGr°/fl7+ln Q=ln Q/K
where AGr° is the standard state free energy change of the
reaction, R is the gas constant, Tis temperature in degrees
Kelvin, Q is the reaction quotient (or ion activity product),
and Kr is the temperature- and pressure-dependent equilib-
rium constant of a reaction. Another term used frequently
in place of the saturation index is the relative saturation,
Q = Q/Kr. At chemical equilibrium, AGr°= 0, Q= Kr, and
Q = 1. In this special case, the solution of interest is in
equilibrium with the mineral and no dissolution or precipita-
tion should take place. Where AGr°< 0, the mineral cannot
precipitate from solution and the thermodynamic driving
force is such that mineral dissolution should occur. Where
AGr°>0, the mineral will likely precipitate if there are no
limiting kinetic factors (Table 2.3).
Table 2.2 Major mineral classes in aquifers and soils.
Mineral Class
Primary Mineral
Contaminant Precipitate
Hydroxides
AI(OH)3, gibbsite
Fe(OH)3, hydrous ferric oxide
FeO(OH), goethite
FeO(OH), lepidocrocite
Cu(OH)2
Cr(OH)3
Zn(OH)2
Oxides
Fe3O4, magnetite
Fe2O3, hematite
MnO2, pyrolusite
SiO2, quartz
UO uraninite
Carbonates
CaCO3, calcite/aragonite
FeCO3, siderite
MnCO3, rhodochrosite
CdCO3, otavite
ZnCO,, smithsonite
o
PbCO,, cerussite
o
Sulfates
BaSO4, barite
CaSO42H2O, gypsum
PbSO4, anglesite
RaSO,
Sulfides
FeS, mackinawite
FeS2, pyrite/marcasite
PbS, galena
NiS, millerite
HgS, cinnabar
ZnS, sphalerite
Phyllosilicates
AI4(OH)8Si4010, kaolinite
K15AI2(OH)2Si25AI15010jllite
Ni3Si2O5(OH)4, nepouite
Na03Zn3(Si,AI)4O10(OH)2 4H2O, sauconite
34
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The ion activity product is a useful probe to evaluate the
potential for contaminant precipitation. Some caution,
however, is recommended in interpreting solution indicators
as evidence for the presence of a particular precipitated
solid within the plume. The observation that an ion activity
product is equal to a corresponding solubility product is not
unequivocal evidence that a given phase is at equilibrium
or even present in the system (Sposito, 1984). Similarly,
an ion activity product that is greater than a corresponding
solubility product cannot be taken as confirmation that pre-
cipitation is occurring. To understand the state of a system
with respect to precipitation and dissolution, it is recom-
mended that the presence of the relevant solid phases be
evaluated in addition to measuring the concentrations of
solutes that participate in a precipitation reaction suspected
to occur within the ground-water plume.
Table 2.3 Relationships among Q, K, and Q.
Process
Mineral
dissolution
Mineral
precipitation
Equilibrium
Saturation
Index,
log (Q/IC)
Negative
Positive
0
Relative
Saturation,
Q
<1
>1
1
Q, Kr
QKr
Q = K
IIB.2.2 Coprecipitation
Contaminant plumes are often characterized by concentra-
tions of dissolved solids in excess of that found in ambient
ground water. These elevated dissolved solids may be
derived as a component of the contaminant source or due
to the dissolution of soil or aquifer solids during plume
transport. Examples of these processes include interac-
tions of acid wastes with aquifer solids leading to dissolution
of aquifer minerals (e.g., carbonates or oxyhydroxides) or
the development of reducing conditions driven by microbial
degradation of organic contaminants that result in reductive
dissolution of iron-bearing minerals. With downgradient
transport, changes in ground-water chemistry or interaction
with unimpacted aquifer solids may lead to precipitation
of these major ground-water constituents out of solution.
Contaminants may be removed from ground water at the
location where precipitation of these major ground-water
constituents occurs. This process is called coprecipitation,
since the contaminant is sequestered within a newly formed
precipitate, but only as a trace structural component within
the precipitate. Examples of major precipitate classes with
a coprecipitated contaminant include oxyhydroxides (e.g.
Fe^Cr^OH) ), carbonates (e.g., Ca^Cd^O.^), sulfides
(e.g., Fe1xNixS ), and phyllosilicates. The contaminant
may be coprecipitated in a cationic or anionic form de-
pending on ground-water chemistry and the nature of the
precipitating phase.
For coprecipitates (or solid solutions) the concentration of
the contaminant in ground water in contact with a precipitate
may be reduced significantly below that observed for ground
water in which the concentration of the contact is governed
by the solubility of a precipitate in which the contaminant is
a major structural component (e.g., Ca1 xCuxCO copre-
cipitate vs. CuCO pure precipitate). For example, the
partial molal Gibbs Tree energy of a binary mixture can be
expressed as the sum of two components: a mechanical
mixing term and a free energy of mixing term (A.Gm!xture):
^
+ AG
m!xture
where X1 and X2 are the mole fractions of two components
in a binary mixture. The AGmixture term contains an ideal
component that depends on X1 and X2 and a non-ideal
component dependent on Xr X2, and activity coefficients
in the solid phase (y1 and y2):
AGmMure=fi7-[X1lnX1+X2lnX2] +
For an ideal solid solution, y1 = y2 = 1, so that
R7[X, Iny^+XJnyJ = 0. Ideal mixing may be approached
where the amount of substitution is very low (a dilute solid
solution) or where the mixing cations are closely matched
in size and charge. In this case, the AGm!xture function is a
symmetrical parabola having a minimum at X, = X2 = 0.5.
Hence as a general rule, the free energy of binary mixtures
is less than that of the pure, end-member components. It
follows that the solubility of an ion can be lowered in a
mixed ionic compound relative to the solubility of the pure
compound.
Remobilization of a coprecipitated contaminant will be gov-
erned by the overall stability of the host precipitate, which
may be controlled by ground water parameters such as pH
and/or redox state. In most cases, the identification of a
coprecipitation process cannot be made with a single line
of evidence. Observations of decreased contaminant con-
centrations concurrent with decreases in the concentrations
of major ground-water constituents such as Ca, Fe, or dis-
solved sulfide may be indicative of a coprecipitation event.
It is recommended that this evidence be supplemented with
solid phase characterization approaches, such as chemical
extractions or microanalytical techniques, to confirm that
coprecipitation is an attenuation mechanism.
IIB.2.3 Surface Precipitation
Surface precipitation may result when adsorption leads
to high sorbate coverage at the mineral-water interface.
Surface precipitation can be thought of as an intermediate
stage between surface complexation and bulk precipitation
of the sorbing ion in solution (Farley et al., 1985; Ford et
al., 2001). At low concentrations of the sorbing metal at
the mineral surface, surface complexation is the dominant
process. As the concentration of the sorbate increases,
the surface complexation concentration increases to the
point where nucleation and growth of a surface precipitate
occurs. Surface precipitation can be viewed as a special
35
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case of coprecipitation where the mineral interface is a
mixing zone for ions of surface precipitate and those of the
underlying substrate. It is generally believed that surface
precipitation can occur from solutions that would appear to
be undersaturated relative to precipitate formation based
on considering solution saturation indices. The reason for
this may be due in part to different equilibrium constants
for surface precipitation versus precipitation from solution,
or may be related to the way the mineral-water interface is
modeled (Sverjensky, 2003). For example, the dielectric
constant of water and therefore activity coefficients in bulk
solution may be different from activity coefficients near
a mineral surface. Again, the identification of a surface
precipitation process cannot be made only with solution
data. Solid phase characterization data, such as chemi-
cal extractions or microanalytical techniques, are needed
to confirm that surface precipitation plays a role as an at-
tenuation mechanism.
IIB.2.4 Mineral Transformation
In many cases the solids that precipitate in near surface
environments are not the most thermodynamically stable
phases. For example, hydrous ferric oxide, ferrihydrite,
is metastable relative to the iron oxyhydroxide goethite.
The preponderance of metastability in near surface envi-
ronments is a consequence of the slowness of chemical
reactions at temperatures typical of surficial environments.
Kinetics, therefore, play an integral role in ground-water and
soil geochemistry. Mineral transformation is one example
of how metastable precipitates evolve toward more stable
mineral phases within an aquifer. Ultimately, contaminants
that are initially adsorbed onto or coprecipitated with these
metastable precipitates are likely to become more resistant
to remobilization if they are incorporated into the more
stable transformation product.
The Ostwald Step Rule is often obeyed in low-temperature
mineral formation. Precipitation of less stable and
more soluble phases is followed by transformation to
progressively more stable and less soluble phases. This
behavior stems from the preferential formation of materials
with fast precipitation kinetics over nucleation and growth
of phases with slow kinetics (Stumm, 1992). Differences in
precipitation kinetics are often tied to structural complexity
of the precipitating mineral. Relatively simple structures are
able to form rapidly whereas ordered structures, although
more stable, require longer time periods to develop.
Precursor phases are usually poorly crystalline and they
may be chemically dissimilar to the final stable mineral.
Examples that follow the Ostwald Step Rule include the
precipitation of ferrihydrite and transformation to more stable
iron oxyhydroxides (goethite) and iron oxides (hematite),
the precipitation of mackinawite and transformation to
pyrite, and the precipitation of amorphous aluminosilicates
such as allophane and transformation to halloysite and
kaolinite. Transformation pathways result from solution
mediated processes or solid-phase transitions (Steefel and
van Cappellen, 1990).
The iron monosulfide-to-iron disulfide transformation has
been widely studied in the laboratory and in the field. In this
example of a mineral transformation process, mackinawite
(Fe1+xS) precipitates as concentrations of dissolved sulfide
and ferrous iron accumulate in pore water. It has been de-
termined that in sulfate-reducing environments, pore water
concentrations of ferrous iron and sulfide are controlled
by the solubility of mackinawite. Mackinawite, however, is
metastable with respect to the iron disulfides, pyrite and
marcasite. The rate of transformation from mackinawite
to pyrite or marcasite depends on pH and redox condi-
tions. Metals that coprecipitate with mackinawite are likely
incorporated into pyrite, which is more stable over a wide
pH range and in anoxic conditions. The rate at which this
transformation occurs will be governed by chemical condi-
tions including the coprecipitation or adsorption of contami-
nants and other dissolved constituents from solution. Site
characterization aspects relating to mineral transformation
processes as an immobilization mechanism will involve
determining the spatial concentration distribution of precur-
sor phases and their more stable transformation products
along with contaminant associations using mineralogical
and wet chemical characterization tools.
//B.3 Implications for Natural Attenuation
Assessment
The sorption processes discussed in the preceding para-
graphs may act in isolation or together to arrest contaminant
migration within the aquifer. Factors that dictate which pro-
cess is likely to dominate contaminant attenuation include
chemical properties of the contaminant, chemical charac-
teristics of the ground water, and properties of the aquifer
solids. Due to the complexities of directly identifying the
immobilized form of the contaminant, it is likely that multiple
lines of evidence will be needed to adequately discern the
controlling attenuation reaction. These lines of evidence will
include the evaluation of patterns in ground-water chemistry
that point to potential precipitation or coprecipitation reac-
tions, evaluation of aquifer solids to determine patterns in
contaminant and solid component associations, and the
use of chemical speciation or reaction models to assess
if ground-water and aquifer solid characteristics are con-
sistent with observed contaminant attenuation. Additional
perspective on possible sorption processes for specific con-
taminants is provided in the contaminant-specific chapters
included in Volume 2 and 3 of this document.
IIC. Microbial Impacts on Inorganic
Contaminant Attenuation
The chemical characteristics of ground water and properties
of the aquifer mineral components are, in part, influenced
by microbial reactions. Microbial activity within the aquifer
may also play a more direct role in controlling contaminant
speciation and migration. The influence of microbial reac-
tions may be more pronounced in contaminant plumes that
also contain degradable organic contaminants such as
hydrocarbons or chlorinated solvents. In these instances,
the plume geochemistry may differ significantly from that
observed in ambient ground water at a site. If microbial
reactions play a significant role in contaminant attenuation, it
may be necessary to gather information on the degree that
36
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these reactions control the concentrations and distributions
of reactants that participate in contaminant attenuation,
as well as the capacity of the aquifer to sustain microbial
activity within the plume. In order to provide perspective
on the potential influence of the subsurface microbiology,
the following discussion provides an overview of general
characteristics of subsurface microbiology, the influence of
microbial activity on the redox state within the plume, and
the types of contaminant attenuation reactions that may be
directly mediated by microorganisms.
IIC. 1 Characteristics of Aquifer Microbiology
Microbial reactions can alter the geochemical structure
of the subsurface (Newman and Banfield, 2002), and it
is becoming increasingly important to characterize these
processes in terms of the natural attenuation of inorganic
contaminants. Microorganisms inhabiting the subsurface
environment exhibit a remarkable array of metabolic capa-
bilities enabling them to use organic or inorganic matter as
energy sources and propagate under aerobic or anaerobic
conditions. A large part of a microorganism's metabolism
is devoted to the generation of adenosine triphosphate
(ATP), an "energy currency" that is used by the microor-
ganism for cell synthesis. Microorganisms derive energy
through the oxidation of organic compounds or chemically-
reduced inorganic ions and compounds. The electrons or
hydrogen atoms resulting from oxidation are transferred
in most microorganisms by an electron transport chain to
an electron acceptor which, in the case of aerobic respi-
ration, is oxygen. Some microorganisms are capable of
anaerobic respiration, whereby the electron acceptor is not
O2 but chemically-reducible inorganic compounds such as
NO3", SO42", CO2, Fe3+, or iron/manganese oxyhydroxides.
A third type of metabolism, called "fermentation," involves
intra-molecular oxidation/reduction without an externally
supplied terminal electron acceptor. Fermentation reac-
tions are not considered here because they do not involve
metal transformations.
The metabolic diversity manifested by microorganisms is
key to their being agents of geochemical change in the
subsurface (Ehrlich, 1995). Subsurface biota are usually
categorized as either heterothropic or chemolithotrophic
according to the following criteria:
• Heterotrophic microorganisms - Heterotrophs include
bacteria (single- or multi-celled organisms lacking
internal membrane structures), fungi (mycelial or
single-celled organisms possessing a cell wall but
no photosynthetic capability), protozoa (unicellular,
microscopic animals such as amoebae), and archae-
bacteria (a newly discovered group of organisms pos-
sessing a unique cell envelope, membrane structure
and ribosomal RNA, distinguishing them from all other
microorganisms). Heterotrophs derive energy from the
oxidation of organic compounds and obtain most of their
carbon from these compounds. Some heterotrophs
respire aerobically while others respire anaerobically,
requiring the availability of oxygen or other reducible
compounds to serve as electron acceptors.
• Chemolithotrophic microorganisms - Some genera
of bacteria and archaebacteria are chemolithotrophs.
These microorganisms derive energy from the oxida-
tion of inorganic compounds (for example, Fe2+, HS",
and H2) and obtain their carbon as CO32", HCO3" or
CO2. Some chemolithotrophs are aerobic using O2 as
an electron acceptor while others respire anaerobi-
cally using chemically reducible inorganic matter as
an electron acceptor (Newman, 2001). An example of
this type of metabolism is carried out by methanogens,
which oxidize H2 using CO2 as an electron acceptor to
form CH4. Reactions carried out by chemolithotrophic
microorganisms must derive sufficient energy from
the oxidation/reduction reactions to fix carbon (that is,
reduce CO32~, HCO3~ or CO2 to organic carbon) and
phosphorylate ADR
In the aquifer below the vadose zone the populations and
numbers of microorganisms vary significantly due to limita-
tions to available oxygen. If the aquifer is shallow and is
recharged through cracks and fissures, the water may be
nutrient-rich and oxygenated resulting in an abundance of
different microorganisms whose numbers can reach 106
per gram. Lower population numbers would be expected
in aquifers where the supply of dissolved oxygen from sur-
face recharge is limited. Deep aquifers are generally more
depleted with respect to nutrients (unless contaminated)
and oxygen causing the numbers and diversity of micro-
organisms to decline and often be limited to bacteria and
archaebacteria. This finding has been supported, during
the last 15-20 years, by carefully controlled drilling studies
undertaken to evaluate the microbiology of deep, imperme-
able rock strata (Ghiorse, 1997; Kerr, 2002).
IIC.2 Microbial Controls on Subsurface Redox
State
Since microorganisms chemically transform aquifer con-
stituents such as dissolved oxygen, iron [aqueous and solid
forms of Fe(lll) and Fe(ll)], and sulfur [aqueous species such
as sulfate and dissolved sulfide], their metabolic reactions
may exert significant influence on the redox geochemistry
within the plume. As previously discussed, redox condi-
tions within the aquifer may govern precipitation-dissolu-
tion reactions that control contaminant precipitation or
coprecipitation, as well as the types and concentration of
aquifer solids that may serve as adsorbents. From this
perspective, some knowledge about the microbial popula-
tions as a function of space and time within the plume may
be necessary to understand the existing redox status of the
aquifer and make projections about how it may evolve. For
heterotrophic microorganisms, the electrons or reducing
equivalents (hydrogen or electron-transferring molecules)
produced during degradation of organic compounds must
be transferred to a terminal electron acceptor (TEA). Ob-
servations of microbial systems have led to the development
of a classification system that groups microorganisms into
three categories according to predominant TEAs:
• Aerobic bacteria — Bacteria which can only utilize mo-
lecular oxygen as a TEA. Without molecular oxygen,
these bacteria are not capable of degradation.
37
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• Facultative aerobes/anaerobes — Bacteria which can
utilize molecular oxygen or when oxygen concentrations
are low or nonexistent, may switch to nitrate, manga-
nese oxides or iron oxides as electron acceptors.
• Anaerobes — Bacteria which cannot utilize oxygen
as an electron acceptor and for which oxygen is toxic.
Though members may utilize nitrate or other electron
acceptors, it can be said that they generally utilize
sulfate or carbon dioxide as electron acceptors.
In any environment in which microbial activity occurs, there
is a progression from aerobic to anaerobic conditions (ul-
timately methanogenic) with an associated change in the
redox status of the system. There is a definite sequence
of electron acceptors used in this progression through
distinctly different redox states. The rate, type of active
microbial population, and level of activity under each of
these environments are controlled by several factors.
These include the concentration of the electron accep-
tors, substrates which can be utilized by the bacteria, and
specific microbial populations leading to the progression
of an aquifer from aerobic to methanogenic conditions
(Salanitro, 1993). This results in a loss of organic carbon
and various electron acceptors from the system as well as
a progression in the types and physiological activity of the
indigenous bacteria.
If microbial activity is high, the aquifer environment would
be expected to progress rapidly through these conditions.
The following scenario outlines a general sequence of
events in which aerobic metabolism of preferential carbon
sources would occur first. The carbon source may be from
organic contaminant sources or other more readily degrad-
able carbon which has entered the system previously or
simultaneously with the contamination event.
• Oxygen-Reducing to Nitrate-Reducing Conditions
- Once available oxygen is consumed, active aerobic
populations begin to shift to nitrate respiration. Denitri-
fication will continue until available nitrate is depleted,
or usable carbon sources become limiting.
• Nitrate-Reducing to Manganese-Reducing Conditions
- Once nitrate is depleted, populations which reduce
manganese may dominate. Bacterial metabolism of
substrates utilized by manganese-reducing populations
will continue until the concentration of manganese oxide
becomes limiting.
• Manganese-Reducing to Iron-Reducing Conditions
- When manganese oxide becomes limiting, iron
reduction becomes the predominant reaction mecha-
nism. Available evidence suggests that iron reduction
does not occur until all Mn(IV) oxides are depleted. In
addition, bacterial Mn(IV) respiration appears to be
restricted to areas where sulfate is nearly or completely
absent.
• Iron-Reducing to Sulfate-Reducing Conditions - Iron
reduction continues until substrate or carbon limita-
tions allow sulfate-reducing bacteria to become active.
Sulfate-reducing bacteria then dominate until usable
carbon or sulfate limitations impede their activity.
• Sulfate-Reducing to Methanogenic Conditions - Once
usable carbon or sulfate limitations occur, methano-
genic bacteria are able to dominate.
Development of a general knowledge of the redox status
of the aquifer throughout the plume is important relative to
understanding the processes contaminant attenuation (or
lack thereof) within the plume, as well as changes in the
capacity for or stability of contaminant attenuation with the
return to ambient conditions. There are several approaches
to evaluating the redox state of the aquifer. One common
approach is to monitor the oxidation-reduction potential in
ground water using a platinum electrode with recalculation
of the measured electrode potential as Eh. However, due
to the lack of internal redox equilibrium in natural systems
(Morris and Stumm, 1967) and limitations to electrode po-
tential measurements in a complex matrix such as ground
water, this measurement should be supplemented with other
site-specific determinations. One supplemental approach to
assessing redox status is the determination of concentra-
tions of specific redox sensitive species such as oxygen,
Fe(ll), hydrogen sulfide, or methane as qualitative guides
to the redox status of ground water (Lindberg and Runnells,
1984; USEPA, 2002). These measurements provide a
means to generally characterize the ground water as oxic,
suboxic, or anoxic and may provide more direct evidence of
the occurrence of iron-reducing, sulfate-reducing or metha-
nogenic conditions. Another approach that may be used
to indicate the terminal electron acceptor process (TEAR)
which dominates within the plume is measurement of the
hydrogen (H2) concentration in ground water (Lovely and
Goodwin, 1988). Hydrogen concentration for the various
TEAPs are shown in Table 2.4 (Chapelle et al., 1995).
Table 2.4 Range of hydrogen concentrations for a
given terminal electron-accepting process
that can be used for classification of the
redox status within the contaminant plume.
Terminal Electron
Accepting Process
Denitrification
Iron (III) Reduction
Sulfate Reduction
Methanogenesis
Hydrogen (H2)
Concentration
(nanomoles per liter)
<0.1
0.2 to 0.8
1 to 4
5 to 20
The focus of this discussion has been on the influence of
microbial activity on the redox status of ground water, which
governs the types of aqueous and solid phase reactants that
may be involved in contaminant attenuation. The following
discussion elaborates on the influence of microbial reactions
on the chemical speciation of inorganic contaminants.
38
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//C.3 Impacts on Contaminant Spec/at/on and
Attenuation
Microbe/metal interactions result in immobilization, mobili-
zation, or transformation by extracellular precipitation reac-
tions, intracellular accumulation, oxidation and reduction
reactions, methylation and demethylation, and extracelluar
binding and complexation. Many of these processes play
an important role in achieving natural attenuation through
immobilization or degradation. Many transformations of
metals and metalloids are carried out by subsurface micro-
organisms for the purpose of obtaining energy for growth
and reproduction; however, these transformations may not
result in decreasing the mobility or toxicity of the metal.
Some microbial transformations are carried out specifically
to detoxify their environment, suggesting that microbes have
an enormous capacity to sustain life. Therefore, the role of
microbes in the subsurface is multifaceted. The principal
mechanisms employed by microorganisms to transform
metals and metalloids in the subsurface environment are:
• Oxidation and reduction reactions resulting from mi-
crobial respiratory activities (Lovely, 1993; Ahmann et
al., 1994);
• Biosorption by cell walls, cell membranes and exo-
polymers and intracellular bioaccumulation (Beveridge,
1989); and
• Methylation and demethylation (Oremland et al., 1991:
Bentleyetal.,2002).
These microbial mechanisms are described with emphasis
on reactions involved in natural attenuation of metals and
radionuclides.
IIC.3.1 Contaminant Oxidation-reduction
Reactions
Oxidation-reduction (redox) processes are chemical reac-
tions that involve a transfer of electrons between reactants
and products and consequently a change in the oxidation
state of elements that can either be oxidized to higher va-
lence states or reduced to lower valence states. Changes
in the valance state of a particular element can result in
the following:
• Change in speciation resulting in lower or higher toxic-
ity
• Change in speciation resulting in lower element solu-
bility (immobilization) or increased solubility (mobiliza-
tion)
• Change in speciation that impacts adsorption/desorp-
tion behavior
The contaminants for which redox-mediated processes
may lead to attenuation include elements that can occur
in multiple oxidation states under environmentally rel-
evant conditions. Thus, a transition metal such as Cr is
redox-active since two stable oxidation states (VI and III)
are readily observed in aqueous solution, whereas Ni is
not redox-active as only one oxidation state (II) occurs in
aqueous solutions at ambient conditions. (More detailed
discussions on the redox activity of individual contaminants
are provided in Volume 2.)
Oxidation-reduction (redox) reactions, which can be carried
out by both heterotrophic and chemolithotrophic microor-
ganisms, are either energy-generating or enzymatically-
catalyzed without coupling to energy generation. Microbi-
ally catalyzed, energy generating redox reactions tend to
be rapid and the metal transformations so extensive that
the geological influence is often substantial. One example
that results in the attenuation of metals is the reduction of
soluble SO42' by Desulfovibrio and Desulfotomaculum spe-
cies involving oxidation of organic matter or H2 for energy
and use of sulfate as a terminal electron acceptor (Lovely,
1993). This redox reaction yields dissolved sulfide that
can chemically react with a number of soluble, divalent
metals producing insoluble metal sulfides. In some cases,
the microbially-generated sulfide may chemically reduce
an inorganic contaminant. For example, dissolved sulfide
may result in reduction of arsenate to arsenite. Examples
of enzymatically-catalyzed redox reactions that immobilize
metals for attenuation in the subsurface are the reduction
of mobile Cr(VI) to the less toxic and less mobile Cr(lll)
(Palmer and Wittbrodt, 1991) and reduction by some
microorganisms of Se(VI) or Se(IV) to insoluble Se(0)
(Ehrlich, 1995).
IIC.3.2 Biosorption and Intracellular
Bioaccumulation
Biosorption, a form of contaminant adsorption, is the binding
of inorganic ions to the outer surface of a microorganism by
ion exchange, van der Waals attractions, or chemical reac-
tions between the metal ion and cell wall, cell membrane,
excreted metabolites, or exopolymers. Biosorption is not
dependent on metabolism by the microbial cell. Intracel-
lular bioaccumulation, or absorption, is the transport of
metals across the cell wall and membrane and subsequent
retention of the metals within the cell. Bioaccumlation is
metabolism dependent because the transport process
requires energy.
The outer surfaces of microorganisms are chemically
complex with each group of organisms possessing differing
cell wall, membrane, and expolymeric structures. These
structures are composed of macromolecules having
functional groups including carboxylates, amines, imidazole,
phosphate, sulfydryl, sulfate, hydroxyl and others. Usually
the microbes exhibit a net negative charge owing to the
abundance of carboxylate and phosphate groups, however,
positively charged amines and imidazole functional groups
impart polyfunctional characteristics to the cell surface.
Most inorganic contaminant binding occurs after initial
complexation and neutralization of the chemically active
site. This stoichiometric interaction between contaminant
ions and active sites within the wall acts as a nucleation
site for the deposition of additional ions from solution in a
manner similar to surface precipitation onto aquifer minerals.
The deposition product then grows in size within the
intermolecular spaces of the wall fabric until it is physically
39
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constrained by the polymeric meshwork of the wall.
Therefore, the amount of contaminants bound to the outer
surfaces of organisms can be considerably greater than that
accounted for by the number of charges available because
binding occurs not only by ion exchange type reactions
and covalent bonding to charged functional groups, but by
precipitation and reduction reactions (Beveridge, 1989).
As with mineral surfaces in the aquifer solids, pH plays a
critical role in the binding of inorganic contaminants. At
low pH values cationic metal binding is diminished because
negative charges on functional groups become protonated,
reducing the negative charge density of the organism's sur-
face. Although the quantitative role of extracellular binding
in controlling the concentration of inorganic contaminants
in the subsurface is not well defined, it does play a larger
part in metal mobility than intracellular metal accumulation.
The relative contribution of microbial binding processes to
the overall immobilization of the contaminant within the
plume is largely unknown. However, a simple comparison
of the relative mass of microbes versus reactive minerals
in aquifer solids would suggest that microbial binding plays
a minor role.
//C.3.3 Methylation and Demethylation
Microbial interactions with some inorganic contaminants
considered in this document result in the transfer of methyl
groups onto these elements by a process known as meth-
ylation. This process may result in either the increased
toxicity or mobility of some inorganic contaminants. Bio-
methylation of metal(loid)s takes place when suitable or-
ganisms are present under anaerobic conditions with high
concentration of available metal(loid)s and methyl donors.
These conditions may exist in contaminated ground water
(e.g., derived from landfill leachate). An example of this
type of process is the methylation of inorganic selenium
to dimethylselenide, a volatile product. Methylation of
selenium decreases rather than increases its toxicity and
probably represents a detoxification reaction that occurs
readily in aerobic soils and sediments. Arsenate can also
be methylated by some bacteria and filamentous fungi.
Biosynthesis of organic compounds containing arsenic such
as arsenocholine, arsenobetaine, and arsenosugars also
occurs among various bacteria, algae, and many higher
organisms (Andreae and Klumpp, 1979).
IIC.4 Implications for Natural Attenuation
Assessment
Based on the previous discussion, it is apparent that the
subsurface microbial community could influence inorganic
contaminant transport. For many of the contaminants dis-
cussed in this document, microbial reactions will primarily
exert an indirect influence on contaminant attenuation by
governing the types, distribution, and concentrations of
aqueous and solid phase reactants. However, microbial
activity may constitute a more direct impact on contaminant
speciation, including changes in the redox state of contami-
nants such as arsenic or selenium. From this perspective,
it is recommended that the site characterization effort incor-
porate analysis of the extent to which microbial processes
may govern ground-water chemistry within the plume. This
is more critical at sites where the co-occurrence of readily
degradable organic contaminants result in ground-water
conditions significantly different than those observed in
ambient ground water. For these situations, information on
the mechanism, rate, capacity, and stability of contaminant
attenuation may need to be evaluated within the context of
the behavior of the organic component of the plume.
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Section III
Site Characterization to Support Evaluation of MNA
Richard T. Wilkin, Steven Acree, Steve Mangion, Patrick V. Brady, James E. Amonette, Paul
M. Bertsch, Douglas B. Kent, Robert W. Puls, Ann Azadpour-Keeley, Robert G. Ford
In determining whether MNA is applicable to a site, char-
acterization data are collected to define site hydrogeology
and the reaction(s) that result in contaminant attenua-
tion. Characterization of attenuation reactions will involve
identification of the reactants that control contaminant
immobilization or degradation. The reactants may be of
aqueous and/or solids forms, with solid reactants potentially
consisting of aquifer solids or microbes. Once the specific
reaction mechanism is identified, site characterization will
then proceed to determination of the capacity of ground-
water conditions to sustain the attenuation reaction and
assessment of the stability of immobilized contaminants
to resist remobilization due to changes in ground-water
chemistry. The types of samples collected from within
the plume boundary will include ground water and aquifer
solids. Analyses of these samples will include those more
commonly employed to understand the overall geochemical
context of subsurface chemistry, as well as more specialized
analyses used to identify the chemical speciation of the
contaminant in dissolved or solid form. Simple observa-
tions of the distribution of the contaminant between solid
and liquid forms will be insufficient to identify the specific
reaction process. A mechanistic understanding of the reac-
tion is needed in order to assess the overall performance
characteristics of the attenuation process. This section
will provide an overview of the analytical techniques that
are used to provide evidence for processes and chemical
speciation in the subsurface.
IMA Site Hydrogeology
Hydrogeology is the foundation of the conceptual model
for natural attenuation (National Research Council, 2000).
Three-dimensional characterization of the ground-water
flow field and the changes that occur through time are
crucial to understanding the transport and, ultimately, the
fate of contaminants. Discussions of hydrogeologic site
characterization in this document draw freely from USEPA
(2003) and are largely limited to evaluations performed in
saturated porous-media settings, such as unconsolidated
aquifer materials. Although similar concepts are sometimes
applicable in other geologic settings, such as fractured rock
formations, ground water often moves primarily through
discrete pathways at these sites. In such highly anisotropic
settings, the direction of ground-water flow and contaminant
transport often cannot be determined within a sufficient
degree of certainty to support assessments of natural at-
tenuation rates or processes. Assessment of contaminant
transport and fate in these and other highly heterogeneous
settings is an area of continuing research and is beyond the
scope of this discussion. In similar fashion, hydrogeologic
characterization of vadose zone materials is not specifically
addressed in this section. Flow conditions are often more
complex in the vadose zone than in similar saturated media.
However, characterization of the hydrogeologic properties
of the vadose zone, as well as the saturated zone, will be
necessary at some sites. A general discussion of vadose
zone characterization for contaminant transport assess-
ments is found in USEPA (1991; 1993a; 1993b).
IIIA.1 Characterization Objectives
Much of the spatial variability in observed contaminant
concentrations is the result of geologic heterogeneity. Dif-
ferences in the geologic properties of aquifer materials from
the micro-scale (e.g., grain-size distribution) to the field
scale (e.g., stratigraphy) result in differences in hydraulic
conductivity and the directions and rates of ground-water
flow. Three-dimensional hydraulic gradients that occur due
to the site-specific characteristics of ground-water recharge
and discharge are also dominant controls on flow. Together,
the differences in the hydraulic conductivity of different
aquifer materials, hydraulic gradients, and the changes in
gradients that often occur in response to natural or anthro-
pogenic changes in ground-water discharge and recharge
control the directions and rates of ground-water flow.
Detailed hydrogeologic characterization is essential for
evaluating natural attenuation processes and rates, as
well as for specification of the locations and frequencies
for performance monitoring. The first step in character-
izing contaminant transport, and, therefore the processes
that may result in attenuation, is the determination of the
path or paths of ground-water transport along which the
contaminant will migrate. At a minimum, the hydrogeologic
database should be sufficient to:
• Define geologic and hydrologic controls on the ground-
water flow field (e.g., transmissive units, preferential
pathways, and barriers to flow);
• Quantify flow rates and directions and their spatial and
temporal variations;
• Support characterization of contaminant sources, the
distribution of contaminants of concern, and the effects
of dominant natural attenuation processes; and
• Identify possible receptors.
43
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The scale and intensity of this characterization is deter-
mined by site-specific conditions (e.g., variability in geology
and hydrology and the acceptable level of uncertainty in
the outcome of the evaluations).
IIIA.2 Geologic Characterization
IIIA.2.1 Saturated Porous Media
The key geologic factors that control ground-water flow at
sites where secondary porosity features such as fractures,
including faults, joints, and partings along bedding planes,
are not present include site stratigraphy and lithologic
characteristics. Stratigraphy is the science concerned with
the form, composition, thickness, areal extent, sequence,
and correlation of both consolidated and unconsolidated
geologic materials. It defines the framework of the ground-
water flow field and constrains the pathways for contaminant
migration (USEPA, 1991). Lithology is the description of
the physical and mineralogic characteristics of geologic
materials. Physical characteristics that control flow include
grain-size distribution, grain shape, and packing. As dis-
cussed in other sections, mineral composition and coatings
affect inorganic constituent distribution through a variety of
chemical reactions including cation exchange, substitution,
precipitation, dissolution, acid-base reactions, complexation
and redox reactions (USEPA, 1991). Grain-size distribution,
grain shape, and packing (i.e., arrangement) of grains influ-
ence the hydraulic conductivity, total porosity, and effective
porosity of the geologic materials.
At many sites, evaluation of sedimentary depositional
environments is an especially useful framework for the
understanding of site stratigraphy and the distribution of
lithologic controls on ground-water flow. Sedimentary facies
(i.e., sedimentary bodies that are internally similar in char-
acteristics) determine the three-dimensional geometries,
connectivity, and heterogeneity of aquifer units and barriers
to flow (Galloway and Sharp, 1998) in many porous-media
settings. Ground water can move preferentially through
coarser-grained materials, resulting in varying degrees of
heterogeneity in flow patterns and contaminant transport.
Interconnected facies of relatively coarse-grained materi-
als may provide preferential pathways for contaminant
migration. An example of a naturally occurring preferential
pathway would be a deposit of interconnected sands of high
hydraulic conductivity bounded by deposits of finer-grained
materials. In similar fashion, anthropogenic features such
as buried utility corridors and heterogeneous fill materials
may also result in the formation of preferential pathways for
ground-water flow or impediments to flow. However, even
within a given preferential pathway, ground water may move
in sinuous paths due, in part, to small-scale differences
in hydraulic conductivity of the materials or to temporally
variable, three-dimensional hydraulic gradients.
Interconnected transmissive materials may be separate and
distinct pathways for contaminant movement. For example,
the degree of hydrologic connection between different sedi-
mentary facies depicted in Figure 3.1 is small. Monitoring
points in different facies may appear to be similar and may
be hydraulically down gradient of one another without sig-
nificant ground-water flow and contaminant transport from
one unit to the next. This is explicitly illustrated in Figure
3.1 where two wells are screened at similar depth below the
water table along the pathway of ground-water transport.
The two screens are placed within two different facies that
likely possess different magnitudes of hydraulic conductivity
(i.e., Well 1 screened in a higher conductivity material in
the "medium to coarse-grained sand" vs. Well 2 screened
in a lower conductivity in the "fine-grained silty sand"). The
difference in hydraulic conductivity between these two fa-
cies may prevent a direct line of transport from Well 1 to
Well 2. In such cases, apparent contaminant attenuation
between monitoring points (i.e., due to lower concentration
observed in Well 2) may be an artifact of sample location
and not representative of actual conditions. Inferences
about natural attenuation based on apparent decreases in
contaminant concentration in the down gradient direction
are likely to be incorrect in these situations unless ground-
water flow paths are determined and monitored. This level
of characterization is often difficult to accomplish using
small numbers of monitoring points.
Three-dimensional characterization will be needed to evalu-
ate and predict the effects of natural attenuation processes
(e.g., advection, dispersion, diffusion, and sorption pro-
cesses) on contaminants at many sites. Data required to
construct the site geologic framework include stratigraphic
and lithologic data obtained from geologic cores and sup-
plemented with information from surface and, particularly,
borehole geophysical methods. Innovative characterization
technologies, such as the cone penetrometer and geologic
sampling using direct-push methods, offer the potential for
cost-effectively evaluating the geologic controls on ground-
water flow and their variability in greater detail than previ-
ously possible without obtaining continuous cores during
traditional drilling. Regardless of the chosen technology,
the choice of sampling locations needed as a basis for
evaluations of natural attenuation processes depends
on factors such as the site-specific geology, contaminant
characteristics, degree of physical, chemical, and biological
heterogeneity, and the locations of points of compliance
and or critical down gradient exposure points.
IIIA.2.2 Saturated Fractured Media
In rock or consolidated materials, features such as frac-
tures often control ground-water flow. The primary factors
affecting flow through fractured media are the density and
orientations of fractures, the effective aperture width of the
fractures, and properties of the rock matrix (Schmelling and
Ross, 1989). An adequate characterization of a fractured
rock system would generally include information on fracture
orientation (i.e., strike and dip), aperture widths, fracture lo-
cations, and interconnection; hydraulic head throughout the
contaminated volume; the distribution of rock porosity and
permeability; sources of ground water and contaminants;
contaminant distribution; and chemical interactions with
the rock matrix. Chemical weathering of the rock and the
subsequent mineralogic changes that occur along fractures
can also be especially important controls on the fate and
transport of inorganic contaminants.
44
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I
Contaminant I
Plume
Legend
Gravel,gravel-sand mixtures
Medium to coarse-grained sand
.... Contaminant Plume
Fine-grained silty sand
Bedrock
Figure 3.1 Geologic block diagram and cross section depicting a stream environment. In this setting, numerous
fades of channel materials are surrounded by finer-grained materials of lower hydraulic conductivity
(modified from USEPA, 2003). Two monitoring wells, labeled [1] and [2], are shown in red. A magnified
portion of the aquifer cross-section is shown to the right to clarify that the well screens are placed within
two different fades possessing differing hydraulic conductivites.
Many important features, such as fracture patterns, can
often be determined by standard surficial geologic mapping
techniques. Much additional data concerning subsurface
conditions can also be obtained using downhole and surface
geophysical methods. However, sufficient characterization
to support natural attenuation process evaluations with the
level of certainty needed to satisfy data quality objectives
will be neither technically nor economically feasible at the
scale of many sites.
IIIA.3 Hydrologic Characterization
Knowledge of the geologic framework provides much infor-
mation needed to define the features that control ground-
water flow. However, data regarding the hydraulic properties
(e.g., hydraulic conductivities and effective porosities) of the
aquifer materials and their distribution are needed to allow
quantitative estimates of such parameters as ground-water
flow rates, contaminant fluxes, and contaminant attenua-
tion rates. Much of the necessary data regarding hydraulic
properties will generally be acquired through hydraulic tests
performed in the field (e.g., pumping tests, slug tests, packer
tests, and tracer tests). Additional information regarding
geologic and hydrologic site characterization concepts and
techniques may be obtained from a variety of sources (e.g.,
Butler, 1998; Kruseman and de Bidder, 1989; USEPA, 1991;
USEPA, 1993a; USEPA, 1998).
Hydrologic characterization of the ground-water flow
system also requires an understanding of natural and
anthropogenic sources for recharge, the characteristics of
discharge, and the hydraulic gradients that are the result.
Recharge sources include precipitation, surface-water
bodies, irrigation, and losses from potable water distribution
systems. Important locations of ground-water discharge
include water production wells, springs, wetlands, and
surface-water bodies. The locations and rates of ground-
water recharge and discharge are important factors in
determining site-specific hydraulic gradients and, therefore,
the directions and rates of flow. Hydraulic gradients may be
three dimensional in nature with strong vertical as well as
horizontal components in many cases. This often results
in a dynamic, three-dimensional flow system.
45
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Temporal variations in either natural or anthropogenic
recharge or discharge may result in fluctuations in both the
horizontal and vertical components of hydraulic gradients,
the directions/rates of contaminant migration, contaminant
loading to ground water, and, potentially, redox conditions.
These fluctuations may lead to changes in the geometry
of the plume that should be considered during evaluations
of natural attenuation and monitoring plan development.
For example, seasonal changes in precipitation may result
in non-uniform changes in water table elevations due to
differences in recharge related to topography, different soil
types, or land uses. This may result in seasonal changes
in hydraulic gradients or discharge locations. Rapid and
sustained changes in ground-water flow velocity are
common in flood plains, particularly near large rivers that
have major changes in the river stage. In some situations,
changes in contaminant loading to ground water may also
occur due to increased recharge through contaminated
vadose zone materials or elevation of the water table into
these materials. Longer-term patterns associated with
sequences of unusually wet or dry years may also influence
ground-water velocity over correspondingly longer time
periods.
Anthropogenic influences on site hydrology, such as
changes in ground-water withdrawal or irrigation rates and
patterns, may have similar effects on plume behavior but
occur on frequencies other than those corresponding to
precipitation patterns. Irrigation or municipal water sup-
ply wells that pump intermittently can affect ground-water
flow patterns in a complex manner that may be difficult to
assess. In addition, land use changes that alter patterns
of recharge, discharge, or withdrawal may be important
sources of variability in the ground-water flow field that
should be routinely considered during the life of an MNA
remedy. The illustration in Figure 3.2 provides an example
where the off-site installation and activation of an irrigation
well can result in movement of the ground-water plume in
Contaminant Monitoring Well
Concentration _>l"sters
\
\
Ground-Water Flow
Active Irrigation Well o
,*
Ground-Water Flow
Time 1
Time 2
Active Irrigation Well •
New Monitoring Wells
A A
Ground-Water Flow
TimeS
Well Cluster
W3
Well Cluster
W3
(Offset)"
Irrigation
Well
Well Cluster
W3
(Offset)*
Irrigation
Well
* Offset from plane of cross section
W3-B
W4-B
W5-B
Time
Figure 3.2 Potential effects of changes in ground-water flow direction on temporal trends in contaminant concentra-
tions (USEPA, 2003).
46
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response to this new ground-water withdrawal from the
aquifer. In this situation, a decrease in contaminant con-
centration at a well screen originally placed within a portion
of the plume with highest concentrations could be mistaken
for evidence of an attenuation process. This misinterpreta-
tion in contaminant concentration trend can be prevented
through periodic re-examination of the hydraulic gradient
distribution within the monitored portion of the aquifer. In
addition, these observations also provide guidance on how
to adjust the location of monitoring points within the aquifer
to insure that the entire plume is being monitored.
Elevations of surface-water bodies and ground-water el-
evation data periodically obtained from a network of wells
and piezometers screened at appropriate depths within the
contaminated aquifer units and surrounding units are es-
sential elements of hydrologic analyses. Additional data for
assessing ground-water flow include land use information
such as the locations, rates, and schedules of irrigation;
local precipitation data; and pumping rates and schedules
for nearby wells. These data and the resulting estimates of
hydraulic gradients are used in conjunction with interpreta-
tions of the hydraulic properties of subsurface materials
and locations of contaminant sources to evaluate potential
changes in contaminant loading, transport and attenuation
rates, and transport directions with time. The evaluations
should be three-dimensional in nature, including horizontal
and vertical components of hydraulic gradients that result
in three-dimensional contaminant transport.
In many cases, frequent (e.g., weekly or monthly) monitor-
ing of ground-water and surface-water elevations may be
warranted, particularly during early phases of monitoring, to
improve the characterization of ground-water flow patterns.
In some cases, monitoring of these parameters on a very
frequent basis using automated recording equipment may
be needed to determine the effects of variability in recharge
and discharge rates or locations. Once the effects are de-
termined, the information would be used in the specification
of appropriate long-term monitoring frequencies. These
data may also indicate changes in hydraulic gradients that
warrant more frequent monitoring of chemical parameters.
Based on the results of such assessments, monitoring
frequencies may be adjusted to adequately capture the
fundamental features of the observed trends. In general,
several years of monitoring data are often necessary for
estimation of the site variability in the ground-water flow
field.
IIIA.4 Ground-WaterlSurface-Water Interactions
Ground-water/surface-water interactions are of particular
importance at sites where surface-water bodies are present.
The hydrologic, as well as geochemical conditions, in
areas where ground water discharges to or is recharged by
surface water often differ markedly from those in the main
body of the plume and may require intensive monitoring
to determine the effect on remedial goals (Winter, 2000).
Locations where ground water discharges to surface water
may vary both temporally and spatially due to changing
hydrologic conditions (Winter et al., 1998; USEPA, 2005a).
Tools to characterize the hydraulic relationships between
ground water and surface water include piezometers,
devices for the direct measurement of ground-water flux
and velocity, geophysical methods, and certain geochemical
techniques. These tools may be used to define areas of
plume discharge into the surface-water body, quantify rates
of discharge, and to aid in the selection of monitoring point
locations for determining the impact of the discharging
water on the sediments and surface-water quality. Multilevel
monitoring is generally required to characterize the
interactions of ground water and surface water features.
For situations where the contaminant plume intersects the
transition zone between ground water and surface water,
information on the distribution of ground-water discharge/
recharge can be used in the selection of monitoring points to
characterize chemical characteristics of the transition zone.
Characterization of the transition zone can be employed to
assess and demonstrate that contaminant concentrations in
discharging ground water or accumulated in surface water
sediments do not negatively impact human or ecological
receptors. Additional discussions of methods for hydrologic
characterization of ground-water/surface-water interactions
and monitoring of the potential for contaminated ground-
water discharge to surface water may be found in various
sources, including USEPA (2000a).
IIIA.5 Hydrogeologic Data Interpretation
The data obtained during hydrogeologic characterization
are used in conjunction with information on contaminant
sources, distribution, and behavior; redox conditions; and
possible receptors to develop a conceptual model de-
scribing site conditions (Figure 3.3). A three-dimensional
conceptual site model incorporating temporal changes is
generally necessary to provide a framework for interpreting
the site data, judging the significance of changes in site
conditions, and predicting the range of future behavior of
the source and plume. Understanding plume formation
and behavior is the basis for evaluating whether an MNA
remedy may be able to achieve site remedial goals within
given time frames. Conceptual site models are expressed
tangibly in text, maps (e.g., chemical isoconcentration maps
and potentiometric surface maps), cross sections (e.g.,
hydrogeologic and chemical distributions), plots of temporal
trends, and other graphical formats, and should be formu-
lated in terms of mathematical calculations describing the
plume and site. The conceptual model is a dynamic tool
that is continually challenged, evaluated, and refined as
new data are obtained. The data and analyses necessary
for formulation of an adequate three-dimensional concep-
tual model for describing the effects of natural attenuation
processes depend on site-specific conditions.
The development of quantitative models (i.e., mathematical
models) based on the conceptual site model is often an
important part of site characterization and remedy selection
for MNA. These quantitative models may be as simple as
equations for estimation of ground-water flow rates (e.g.,
Darcy's law), or as complex as numerical models of ground-
water flow and contaminant fate and transport. Such
calculations are used to help understand site processes,
locate monitoring points, estimate attenuation rates, and
evaluate possible effects of different conditions on plume
47
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behavior. Quantitative models require particular types of
data, and the data collection effort should be designed
considering the requirements of the chosen model(s) and
site-specific characterization objectives.
The conceptual site model for natural attenuation is the
site-specific qualitative and quantitative description of the
transport and fate of contaminants with respect to receptors
and the geologic, hydrologic, biologic, and geochemical
factors that control contaminant distribution (USEPA, 2003).
The model expresses the understanding of the structure,
processes, and factors that affect plume development and
behavior. It is built on assumptions and hypotheses that
have been tested using site-specific data.
IIIA.5.1 Attenuation Rate Estimates
The spatial and temporal distribution of the contaminant
within the ground-water plume will depend on source loca-
tion, the spatial distribution and velocity of water flow (or
diffusion where advective flux is slow), and the abundance
and biotic/abiotic reactivity of aqueous and solid phase
biogeochemical components along the paths of water flow.
The extent to which biogeochemical processes will cause
contaminant attenuation during transport will depend on the
relative rates of both fluid transport and chemical reactions
(Morgan and Stone, 1985). For example, contamination
will likely be negligible for systems in which the timescale
for water flow within a hypothetical reaction volume is much
shorter than the timeframe for significant reaction to take
place. Conversely, significant attenuation may observed
for contaminants in systems where reactions occur rap-
idly relative to the timescale for fluid transport. The latter
situation is commonly assumed to apply to ground-water
systems, but this needs to be confirmed during site char-
acterization. This is particularly important for near-surface
systems that may experience large variability in water flux
and fluid velocities (e.g., Conant Jr., 2004).
Attenuation rate constant calculations can be an important
tool for evaluating the feasibility of natural attenuation at
a contaminated site, e.g., as part of the Tier II screen-
ing analysis. Specific applications identified in U.S. EPA
guidelines (USEPA, 1999) include use in characterization
of plume trends, as well as estimation of the time required
for achieving remediation goals. As illustrated by Newell
et al. (2002) different types of attenuation rate data may be
obtained for a given site and careful consideration should be
given to the appropriate use of the first-order rate constants
derived from these data.
The illustrations in Figure 3.4 provide two examples of
attenuation rate constant data that may be collected within
the plume. The overall extent of plume attenuation can
be estimated by examining contaminant concentrations
at a series of wells within the ground-water flow system
(Panel A). In this instance, the influence of dispersion on
contaminant concentrations as a function of distance must
be determined through comparison with a conservative
dissolved tracer. The duration of a plume at a given
location can be estimated through analysis of contaminant
concentration trends with time (Panel B). This can be
used to estimate the time required to reach ground-water
remediation goals. In both cases, the overall determination
of whether the ground-water plume is expanding, showing
relatively little change, or shrinking (in three dimensions)
will depend on the analysis of multiple well transects
(concentration vs. distance) or well points (concentration
vs. time) in order to provide appropriate spatial coverage.
Since the directions and rates of ground-water flow can
vary in time, the determination of concentration-based
rate constants may be subject to errors of a magnitude
determined to be unacceptable for site characterization.
In order to reduce this level of potential error, one may
determine mass-based rate constants provided details on
the spatial and temporal variability of hydraulic flux is also
well characterized. Newell et al. (2002) should be consulted
for further details on the application and limitations of the
various approaches for estimation of rate constants.
IIIA.5.2 Contaminant Flux
While contaminant concentration is a determining factor
for human or ecological risk, this metric does not provide
a measure of contaminant mass and distribution within the
system of interest. Determination of a contaminant mass
balance is critical for determining changes in contaminant
mass or speciation during subsurface transport. For organic
contaminants, a mass balance calculation aids determi-
nation of whether contaminant degradation or sorption is
occurring. For inorganic contaminants, a mass balance
is required to assess changes in chemical speciation and
mobility. The mass balance is inclusive of liquid and solid
matrices (e.g., aquifer sediments) within the boundaries
of the conceptual model. The distribution of contaminant
mass is important with respect to defining remedial objec-
tives and assessing remedial performance.
Contaminant flux (M) is defined as the product of con-
taminant concentration (C) in the mobile phase (water
and mobile colloids) and the volumetric flow of the mobile
phase (Q):
M = C*Q
Contaminant flux can be calculated for point locations or
cross-sectional areas perpendicular to water flux depending
on the level of heterogeneity in water flow or contaminant
concentration distribution (Einarson and Mackay, 2001;
Buscheck et al., 2004). For inorganic contaminants, this
general equation represents the flux of all contaminant
species at a given point in time. Since a mobile inorganic
contaminant can change chemical form, it may be useful or
necessary to further define contaminant flux for individual
contaminant species relevant to site-specific conditions.
Mi = C, * Q
Contaminant transport occurs along water flow paths in
the subsurface. Therefore, the first step to determining
contaminant flux is developing an understanding of system
hydrology. With the establishment of a water budget for the
site (i.e., water flux distribution), then a mass budget can
be developed to establish contaminant flux across relevant
system boundaries.
48
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Contaminant Source and Source
Control Information
• Location, nature, and history of contaminant releases
or sources
• Locations and characterizations of sources for ground-
water contamination [e.g., nonaqueous phase liquid
(NAPL)]
• Locations and descriptions of source control and other
ongoing and proposed remedial actions
Geologic and Hydrologic Information
Regional and site geologic and hydrologic settings,
including controls on ground-water flow
Analyses of depositional environments and geologic
features that may serve as zones of preferential flow or
barriers to flow, including geometry and physical
properties of geologic facies (eg., texture, porosity, bulk
density) and their variability
Stratigraphy, including thickness and lateral continuity
of geologic units, and bedding features
Anthropogenic features (e.g., buried corridors and
heterogeneous fill materials) that control ground-water
flow, and may serve as migration pathways or barriers
Depth to ground water and temporal variation
Characteristics of surface water bodies (e.g., locations,
depths,and flow rates),their interactions with ground
water, and temporal variations
Ground-water recharge and discharge locations, rates
and temporal variability
Hydraulic gradients, including horizontal and vertical
components, and their variations in response to
fluctuations in site hydrology (e.g., seasonal or longer
term precipitation patterns and changes in patterns of
ground-water withdrawal or irrigation)
Hydraulic properties (e.g., hydraulic conductivities,
storage properties, and effective porosities) and their
variability and anisotropy within geologic units
Quantitative description of the ground-water flow field
Chemical properties of the subsurface matrix including
mineralogy and organic matter
Receptor Information
Aquifer classification, current usage information, and
reasonably anticipated future usage
Locations and production data for water-supply wells
Locations and information on human and ecological
receptors under current and reasonably anticipated
future conditions
Areas susceptible to impact by vapor-phase
contaminants (e.g., indoor air)
Information on local historical and cultural uses of land,
water,and other resources used to identify receptor
populations
Descriptions of institutional controls currently in place
Contaminant Distribution, Transport and Fate
Distribution of each contaminant phase (i.e., gaseous,
aqueous, sorbed, NAPL) and estimates of mass
Mobility of contaminants in each phase
Temporal trends in contaminant mass and
concentrations
Sorption information, including retardation factors,
sorption mechanisms, and controls
Contaminant attenuation processes and rate estimates
Assessment of facilitated transport mechanisms (e.g.,
complexation or colloidal transport)
Geochemical characteristics that affect or are indicative
of contaminant transport and fate, and mineralogy, if
needed
Potential for mobilization of secondary contaminants
(e.g., arsenic)
Effects of other proposed or ongoing remedial
activities on contaminant transport,fate and natural
attenuation processes
Figure 3.3 Elements of a conceptual site model for monitored natural attenuation of inorganic contaminants (modi-
fied from USEPA, 2003).
49
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(A) Determining Plume Attenuation
(B) Determining Plume Duration at a Point
Figure 3.4 Illustration of two approaches for determining attenuation rate constants within a contaminant plume.
Critical assumptions underlie the applicability of these two analysis approaches: 1) analysis depicted in
Panel A presumes that a plume centerline exists and that five monitoring wells are sufficient to define it,
and 2) analysis depicted in Panel B presumes that the plume is stable. Both assumptions are to be veri-
fied through acquisition of supporting spatial and temporal data in order to reduce the uncertainty in rate
estimates.
IIIA.5.3 Source Term Characteristics
In addition to the geologic factors that influence ground-
water transport, the chemical and physical characteristics
of the source term can play a significant role in determin-
ing the physical dimensions of the ground water plume.
The dimensions of the mobile plume may be impacted by
characteristics of the contaminant source such as the total
mass and rate of release of contaminant into the saturated
zone within the aquifer. Active treatment of the source
zone may also impact plume characteristics, particularly
in cases where engineered remedies within the source
zone introduce chemical reactants, stimulate microbial
processes, or influence ground-water flow. Finally, the
chemical characteristics of the contaminant source mate-
rials may impact the contaminant transport either directly
through chemical reactions that influence aqueous specia-
tion (e.g., Pb complexation by EDTA to form solution spe-
cies) or indirectly by influencing ground-water chemistry
(e.g., via biodegradation of organic co-contaminants such
as fuel hydrocarbons or chlorinated solvents). In some
instances, microbial reactions that are stimulated by the
presence of organic contaminants may result in conditions
beneficial to inorganic contaminant attenuation within the
plume, e.g., generation of sulfate-reducing conditions that
result in precipitation of metal sulfides. While this process
may be beneficial to inorganic contaminant immobiliza-
tion, the site investigation needs to include consideration
of the long-term stability of the immobilized contaminant
for a site at which the aquifer is generally oxic outside of
the plume boundaries. This discussion serves to highlight
the importance of factoring in evaluation of contaminant
source characteristics and the management of the source
area relative to down gradient plume behavior.
NIB. Contaminant Quantification, Distribution
and Speciation
Determination of the attenuation mechanism, assessment
of aquifer capacity to attenuate contaminant mass, and the
evaluation of the long-term stability of immobilized contami-
nants will necessitate analytical measurements conducted
on aqueous and solid samples collected from within the
plume. The types of measurements will include determina-
tion of total element concentrations as well as determination
of aqueous and solid phase speciation of the contaminant
and reactants involved in the attenuation reaction. The
measurements will likely be applied to samples collected
over a range of locations and times in order to adequately
characterize the spatial and temporal variability of aquifer
conditions within the plume. The following discussion will
provide an overview of the types of measurements that
may be required to evaluate the adequacy of site-specific
attenuation reactions. The discussion will first address
aqueous phase measurements followed by solid phase
measurement techniques.
IIIB.1 Aqueous Characterization Approaches
Characterization of ground water will include assessment
of the overall chemical conditions as well as the chemical
speciation of the contaminant. Determination of the overall
chemical conditions within the plume (e.g., redox status
discussed in Section NIC) sets the context for evaluating
reactions that lead to contaminant attenuation. Chemical
measurements to assess redox status of ground water can
be achieved through a combination of field- and laboratory-
based methods. Evaluation of the types and distribution
of ions in ground water can adequately be addressed via
laboratory measurements on properly preserved samples.
50
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These data are typically required to assess what type(s)
of sorption reactions control contaminant immobilization.
Likewise, determination of contaminant degradation
products is needed to verify attenuation is occurring. An
underlying assumption for all analysis types is that the
approaches to sample collection and preservation are
adequate to maintain the intact chemical characteristics
of the ground-water sample. Insuring the adequacy of
sample collection protocols can be realized, in part, through
adherence to recommendations for well installation and
low-flow well sampling procedures (USEPA, 1994; USEPA,
1996). It is also recommended that the design of the well
monitoring network consider temporal as well as spatial
variability that may influence chemical measurements.
For example, variations in ground-water levels may result
in variations in ground-water redox chemistry. These
considerations are explored more holistically in USEPA
(2003), which should be consulted both from the perspective
of designing the initial ground-water characterization plan
as well as the long-term performance monitoring plan.
In general, a variety of water samples should be collected
at the same time to allow analysis for metals, common
anions, alkalinity, total dissolved carbon, dissolved oxygen,
and pH. In some systems, assessment of oxidation-reduc-
tion potential (ORP) using a Pt electrode may be desirable
for qualitative assessment of variability between screened
intervals within the plume. For metals and anions, the use
of ultra-clean sample bottles is essential to avoid artifacts,
and these can be checked through the inclusion of trip
blank samples consisting of deionized water or tap water
in the same types of containers.
For preservation, samples for metal analysis are usually
acidified immediately upon collection to pH 2 using either
HNO3 or HCI to avoid precipitation of oxide phases. Sam-
ples for anion analysis are not treated with acid, as this may
swamp the anion of interest during the ion chromatography
analysis. Subsamples of the same sample taken for anion
analysis can be used for alkalinity determinations, as this
quantity is robust. Measurements of dissolved oxygen
and pH are best taken in the field using a closed flow cell
instrumented with electrodes for in-line measurement. For
most samples (especially total dissolved carbon), chilling to
4 °C as soon as possible after collection is useful, although
changes of temperature for dissolved gas measurements
are not recommended. Methods used for determination
of total concentrations of metals in aqueous samples are
usually the same as used for destructive analysis of sol-
ids with elimination of the digestion step. Oxidation-state
determinations can be made on the basis of reactions
with chromophores [e.g., s-diphenyl carbazide for Cr(VI),
phenanthroline orferrozine for Fe(ll)] and determination of
absorbance in the UV-Vis region of the spectrum. Field kits
and ion-selective electrodes are available for many of these
determinations as well as for pH and dissolved oxygen.
Common anions are usually determined in the laboratory
using ion chromatography or capillary electrophoresis.
Colorimetry involves specific complexation of the atom
of interest in solution by a strongly absorbing compound
having a high absorptivity at an accessible wavelength
(i.e., a chromophore) and subsequent measurement of the
absorptivity of the solution. Atomic absorption, emission,
and fluorescence rely on the same electronic transitions
involving valence-shell electrons of unbound atoms. With
absorption, the amount of light lost in exciting electrons
into unbound states is measured, whereas with emission
and fluorescence, the amount of light emitted when these
electrons return to the ground state is measured. Emission
and fluorescence differ in the manner by which the elec-
trons are excited—emission involves thermal excitation of
the sample by injection into a flame (flame photometry) or
a plasma (ICP-AES), whereas fluorescence uses photon
excitation. The sensitivity of colorimetry, AAS, AES, and
AFS are roughly comparable (ppm to ppb), although ICP-
AES has the advantage of being able to analyze multiple
elements more easily with fewer matrix interferences and
across a wider concentration range than the other tech-
niques. With ICP-MS, the masses of individual atoms are
measured using a mass spectrometer, the intake of which
is coupled to a plasma where the atoms are thermally
ionized. This technique has many advantages including a
wide dynamic range, ability to analyze all elements heavier
than He, few matrix or spectral interferences, and the abil-
ity to measure isotopic ratios. Specific recommendations
on contaminant speciation measurements and supporting
ground-water chemistry data are provided within individual
non-radionuclide contaminant chapters included in Volume
2 of this document.
Ill B.1.1 Filtration
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 interpret the 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 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 associated contaminants) that could pass through
0.45 urn filter membranes. The use of filter membranes
with pore sizes of 0.1 um or less will generally provide a
better assessment of the dissolved vs. particulate load of
ground water.
Analytical methods have typically used 0.45 um filters to
differentiate between dissolved and particulate phases. If
the intent of such determinations is an evaluation of truly
dissolved concentrations, which would be important for
geochemical modeling purposes, the inclusion of colloidal
material less than 0.45 um will result in incorrect values.
Conversely, if the purpose of sampling is to estimate 'mobile'
contaminant species in solution, significant underestima-
tions of mobility may result, due to colloidal facilitated trans-
port by particles, which are filtered out by 0.45 um filtration.
Kim et al. (1984) found the majority of the concentrations of
rare earth elements to be associated with colloidal species
that had passed a 0.45 um filter. Wagemann and Brunskill
51
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(1975) found more than a two-fold difference in total iron
and aluminum values between 0.05 and 0.45 urn filters of
the same type. Some aluminum compounds were observed
to pass through a 0.45 urn filter, but were retained on a
0.10 urn filter (Hem and Roberson, 1967). DeVitre et al.
(1987) found approximately 35% of the particulate bound
manganese in the 0.015 to 0.10 um size fraction in anoxic
lake waters. Kennedy and Zellweger (1974) found errors
of an order of magnitude or more in the determination of
dissolved concentrations of aluminum, iron, manganese
and titanium using 0.45 um filtration as an operational
definition for "dissolved". Sources of error were attributed
to filter passage of fine-grained clay particles. De Mora
and Harrison (1983) provided an excellent review on the
subject of physical separation techniques for trace metal
speciation.
During sample collection in anoxic or suboxic systems, iron
oxidation and precipitation may occur prior to filtration and
result in the removal from solution of previously dissolved
species due to instantaneous sorption by the precipitate
(Puls and Eychaner, 1990). Filter loading and clogging of
pores with fine particles may also occur, introducing filtration
errors due to reductions in effective pore size (Danielsson,
1982). Sheldon and Sutcliffe (1969) found that virtually all
filters remove particles smaller than the stated pore size.
In experiments with seawater and latex particles, using
light scattering techniques, Johnson and Wangersky (1985)
demonstrated that a high proportion of materials dispersed
at sizes smaller than the filter pore size will interact with the
filter surface. These interactions are dependent upon size,
particle concentration, colloid surface chemistry, electrolyte
concentration and composition, nature and concentration
of adsorbents, chemical properties of the filter surface, and
the frequency of collisions of dispersed particles with the
filter surfaces.
Contaminants may exist as dissolved species, precipitated
solids, polymeric species or be adsorbed to inorganic or
organic particles of colloidal dimensions. Based on the
above discussions, the filtration of ground water samples
for metal analyses using 0.45 um filters may not provide
accurate data either for geochemical modeling or for con-
taminant migration estimates. Some mobile species are
likely to be removed by 0.45 um filtration prior to chemical
analysis, while other particulate-associated metals will pass
through the filters and be incorrectly considered 'dissolved'.
A principle objective in sampling to test a geochemical
speciation model is to obtain estimates of the free ion ac-
tivities of the major, minor, and trace elements of interest.
Since there are relatively few easily performed analytical
procedures for making these estimates, an alternative is to
test the analytically determined dissolved concentrations
with model predictions, including both free and complexed
species.
Collection of ground-water samples and measurement of
their chemical properties will be necessary irregardless
of whether degradation or attenuation is the primary at-
tenuation mechanism. Characterization of aquifer solids
may be necessary to support evaluation of degradation
processes, but primarily when knowledge of specific mi-
crobial processes or identification of abiotic solid phase
reductants is needed to constrain attenuation rate and
capacity evaluations. In contrast, for contaminants in which
immobilization is the primary (or only) viable attenuation
mechanism, collection and characterization of aquifer solids
is a specific requirement.
IIIB.2 Solid Phase Characterization
Approaches
Determination of aquifer solids mineralogy and solid phase
contaminant speciation is critical to identification of the con-
taminant immobilization process. Contaminant speciation is
inclusive of oxidation state of the immobilized contaminant
as well as the solid phase component(s) with which it is
associated. Determination of contaminant solid phase
speciation can be approached in two ways, structurally and
operationally. Structural determination uses instrumental
procedures (spectroscopy, microscopy) to identify and
quantify the discrete phases present. This information, to-
gether with accumulated knowledge of their thermodynamic
properties and the rates at which they precipitate/adsorb
and dissolve/desorb, is employed to assess the capacity of
the aquifer to sustain contaminant attenuation and evaluate
the long-term stability of immobilized contaminants. The
challenges associated with the structural approach include
1) difficulty in identifying specific phases at concentrations
near or below instrumental detection limits, 2) need for good
sampling statistics (e.g., a large number of micron-sized
measurements are needed to ensure that the site is ad-
equately represented), 3) inadequacy of available thermo-
dynamic and kinetic data to evaluate reaction mechanisms,
and 4) cost of instrumentation and model applications that
may be employed during evaluation of aqueous and solid
phase chemical data.
The operational approach, on the other hand, classifies
contaminant form solely in terms of reactivity. Contami-
nated materials are contacted with solutions that simulate,
in a short period of time, subsurface conditions expected
over much longer time intervals. The amount of contami-
nant released by the material as a result of this contact is
measured. In some instances, the rate of release is also
quantified either by in-situ measurements or successive
extractions. Examples of the operational approach include
chemical extractions using a variety of solutions and pro-
cedures, and bioavailability studies that measure uptake
by organisms directly. Sequential chemical extractions, in
which contaminated materials are treated with successively
harsher solutions, have found wide use in site characteriza-
tion and several standard methods are available for specific
situations (Tessier et al., 1979;Yanase et al., 1991). While
highly relevant, bioavailability studies are rarely conducted
because of the high cost and long time periods required.
Operational determination of contaminant form avoids many
of the problems associated with the structural approach, but
still requires the development of substantial and expensive
databases that correlate the amounts and rates of contami-
nant release during laboratory contact studies with actual
contaminant reactivity under field conditions.
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IllB.2.1 Sampling and Fractionation
Collection of samples is the first step in determination of
contaminant distribution within the source area and down
gradient plume. The most important consideration is that
the samples collected give a fair representation of the
size and extent of the contamination. Enough samples
must be collected to provide statistical certainty to any set
of results. As soils and sediments are heterogeneous,
minimum sample sizes of about 50 g are needed at each
sampling point. These are mixed thoroughly before taking
1- to 10-g subsamples for analysis. In some instances,
mixing of samples from different sampling points to provide
composite samples can be done to decrease analytical
costs, but this practice also tends to decrease the precision
with which the geographical extent of the contamination
can be determined. To avoid the possibility of bias, and
to retain the ability to interpolate results, sampling points
should be laid out on a three-dimensional grid that takes
into account any pre-existing knowledge of the underlying
stratigraphy. Samples may be collected in several rounds,
starting with a coarse grid to identify the overall extent of
the contamination and possible "hot" spots. Successive
sampling rounds using finer grid sizes on subsets of the
original coarse grid can be used to provide further data
about regions of interest.
Aquifer samples can be collected using a variety of methods.
In all cases, however, the goal will be to collect materials
to allow for lithologic logging, to provide representative
samples for laboratory investigations and for submission
to analytical laboratories. The samples must meet the
appropriate data quality objectives as identified in the proj-
ect-specific Quality Assurance Plan (QAP). In all cases,
aquifer materials should be collected with local, State, and
Federal requirements in mind.
The procedures to follow for collection of soil and sediment
samples will depend on the degree to which environmental
availability is to be assessed. Aquifer materials can be col-
lected, for example, using hollow-stem auger or hydraulic
percussion methods. These are generally well-tested
methods applicable to a wide variety of environments.
With these methods various types of sample liners may be
used, such as plastic or brass sleeves. The sleeves can
be cut and capped to preserve materials and facilitate their
transport. If oxidation-reduction processes are believed to
be an important component of attenuation mechanism(s),
special attention must be given to preserving the redox
status of materials after they are retrieved from the subsur-
face. For example, if anoxic materials are collected they
must be frozen after collection or stored in evacuated or
inert-gas-purged containers in order to preserve primary
mineralogy (USEPA, 2002; USEPA, 2006).
For total analyses, little care need betaken and the samples,
once collected, can be air-dried and stored before analysis.
The availability of many metals, however, depends greatly
on their oxidation state, and this may change with time in
storage due to microbial activity or exposure to air. Drying
of samples can also affect the availability of some elements,
e.g., due to irreversible (or slowly reversible) changes in
the structure of illitic and vermiculitic clay minerals during
drying. A realistic assessment of current availability, there-
fore, will attempt to make the measurement as soon after
collection as possible using samples that have been stored
under conditions that maintain their redox and moisture
status at the levels seen in the field (USEPA, 2006). The
best procedures for preserving samples thus involve, at
a minimum, storage of field-moist samples at 4 °C in air-
tight plastic bags from which the air has been squeezed.
In instances where oxidation state is critical, injection of
N2(g) or Ar(g) into the bore hole, and collection of cores
in thick plastic cylinders that are immediately capped and
transferred into anoxic chambers for further processing or
storage is advised. Lowering the temperature to 4 °C slows
microbial activity substantially. Freezing the samples may
result in lysing of microbial cells and can create a different
organic mixture in the soils than was present at the time
of sampling.
In preparation for actual analysis, some particle-size frac-
tionation may be needed for easy sample handling and
enhanced sensitivity. As a result of surface-area consid-
erations, the most reactive fraction of soils and sediments
is that having particle sizes less than about 50 urn (i.e.,
silt and clay). Gravel (particles > 2 mm) and cobbles can
generally be removed by sieving with little or no impact on
the analytical results. The sand fraction (50-2000 urn) is
generally left intact, as its removal by wet-sieving creates
more problems than it solves. Careful records need to be
kept of the mass and volume of the fractions removed so
that analytical results on the remaining fractions can be
scaled to field conditions. Fractionation on the basis of
properties other than particle size can be done to concen-
trate specific minerals, and this may prove beneficial in
some instances (Laird and Dowdy, 1994).
IIIB.2.2 Total Amount
Determination of the total quantity of a metal contaminant
present is generally performed on "bulk" sample sizes of at
most a few grams. In instances where the total amount of
contamination is very low and detection limits are encoun-
tered, analysis by electron or X-ray microscopic techniques
can be performed in which specific particles containing the
contaminant are identified. With the microscopic approach,
sample sizes of ug to mg are common, and scaling to field
concentrations is difficult.
The bulk analysis techniques are distinguished on the basis
of whether they are sample destructive or nondestructive.
The sample-destructive techniques require digestion of the
sample using strong acid or base to destroy the original
compounds present and release the elements in a soluble
state. Once in the soluble state, the concentration of the
contaminant is determined by spectroscopic means after
nebulization or injection into a flame or plasma connected to
a suitable detector. The most common detection techniques
include colorimetry, atomic absorption spectrometry (AAS),
atomic emission spectroscopy (AES) using either flame
(photometry) or an inductively coupled plasma (ICP-AES),
atomic fluorescence spectroscopy (AFS), and inductively
coupled plasma mass spectrometry (ICP-MS).
53
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Colorimetry involves specific complexation of the atom
of interest in solution by a strongly absorbing compound
having a high absorptivity at an accessible wavelength
(i.e., a chromophore) and subsequent measurement of the
absorptivity of the solution. Atomic absorption, emission,
and fluorescence rely on the same electronic transitions
involving valence-shell electrons of unbound atoms. With
absorption, the amount of light lost in exciting electrons
into unbound states is measured, whereas with emission
and fluorescence, the amount of light emitted when these
electrons return to the ground state is measured. Emis-
sion and fluorescence differ in the manner by which the
electrons are excited—emission involves thermal excitation
of the sample by injection into a flame (flame photometry)
or a plasma (ICP-AES), whereas fluorescence uses pho-
ton excitation. The sensitivity of colorimetry, AAS, AES,
and AFS are roughly comparable, although ICP-AES has
the advantage of being able to analyze multiple elements
more easily with fewer matrix interferences and across a
wider concentration range than the other techniques. With
ICP-MS, the masses of individual atoms are measured
using a mass spectrometer, the intake of which is coupled
to a plasma where the atoms are thermally ionized. This
technique has many advantages including a wide dynamic
range (pptr to ppm), ability to analyze all elements heavier
than He, few matrix or spectral interferences, and the ability
to measure isotopic ratios.
The nondestructive analysis techniques generally rely on
inner-shell electronic transitions or nuclear transitions for
elemental identification and thus are less reliant on decom-
position of the sample into individual atoms. Nevertheless, a
homogeneous sample consisting of particle sizes of <2 urn
and presenting a flat surface is needed for best results.
Thus, although the sample is not completely destroyed in
preparation, it is altered by grinding to achieve the particle
size needed. The most common of these techniques is X-
ray fluorescence (XRF), which can use high-energy photons
from an X-ray tube or radioactive source to excite electrons
in the atoms of interest. The light emitted by these atoms
upon return to the ground state is then detected using an
X-ray detector such as a gas-filled proportional counter,
scintillation detector, or solid-state semiconductor detec-
tor. X-ray emission spectroscopy is similar to XRF except
that it uses high-energy particles (electrons, protons, or
alpha particles) to excite the atoms rather than photons.
In general, detection limits are in the ppm range for solids
although X-ray emission spectroscopy is more sensitive
for lighter elements (z<30) and X-ray fluorescence is more
sensitive for heavier elements (z>45) (Amonette and Sand-
ers, 1994).
Microscopic methods of elemental analysis are generally
more expensive than the bulk techniques, but can prove of
great use in identification of specific contaminant phases
and their associations with minerals. Electron microscopy
images particles through scattering of an electron beam. At
the same time, the electron beam is stimulating the emis-
sion of X-rays by the sample and detection of these (as in
X-ray fluorescence) allows simultaneous multielemental
analysis of the material being imaged. Scanning electron
microscopy (SEM) can be used with samples of any thick-
ness, whereas transmission electron microscopy (TEM)
requires preparation of samples on the order of 30-150 nm
thick often in the form of thin sections. Conventional SEM
and TEM require high vacuum for the analysis, due to the
short pathlength of electrons in air, but new developments
of the environmental SEM allows analysis of samples at
pressures approaching ambient thus allowing observation of
wet specimens for short periods of time. With the develop-
ment of synchrotron X-ray sources, a microscopic technique
based on X-ray fluorescence is available for high-value
specimens. X-ray microscopy (XRM) has spatial resolution
near 1 urn, and because of the very high photon fluxes avail-
able can achieve detection limits in the ppb range, better
than any other X-ray technique. The XRM technique can
be coupled with X-ray absorption spectrometry (XAS) to
determine oxidation states and local structural information.
Although primarily a research tool at the moment, XRM
may prove quite useful for phase identification of critical
samples when other techniques fail.
IIIB.2.3 Structurally Defined Form
One approach to the determination of environmental
availability relies on identification of the discrete contami-
nant-bearing phases that are present and combining this
information with phase-relevant thermodynamic or kinetic
data using a geochemical model. Identification and quan-
tification of the discrete phases present is gained through a
combination of elemental, structural, and in some instances,
solubility analysis. Techniques for structural analysis mea-
sure the arrangement of atoms both local to the contaminant
and in extended structures. The combination of structural
and elemental information uniquely determines a thermo-
dynamic phase with specific properties.
The primary technique for structural analysis is X-ray dif-
fraction (XRD), which relies on measuring the coherent
scattering of a collimated beam of X-rays by the sample as a
function of angle to the beam. The technique requires some
degree of repetitive structural unit having dimensions on the
order of the wavelength of the X-rays. Scattering from paral-
lel planes of atoms separated by regular spacings yields an
interference or diffraction pattern. The spacings at which
constructive interference occurs (d) yield high intensity in
the diffraction pattern and can be calculated from the angle
of incidence (6) and the wavelength of the X-ray (k) using
the Bragg equation, d = n?i/2sin9, where n is an integer.
The set of d-spacings and relative intensities measured for
a compound are unique and can be compared with those for
thousands of other compounds stored in a large database.
In soils and sediments, however, mixtures of compounds
occur and overlapping patterns for different compounds
can make identification difficult in many instances. Also,
XRD is sensitive down to about 1 -5% by weight, and thus
is not well-suited for identification of trace phases typical of
contaminants in soils and sediments. To some extent, these
drawbacks can be overcome by fractionating the sample
to eliminate interfering compounds and to concentrate the
contaminant-bearing phases. Diffraction also occurs with
electron beams, as these have wave properties (albeit with
shorter wavelengths than X-rays). Electron diffraction is
used in TEM analysis to identify phases at the same time
54
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they are being imaged and their elemental compositions
determined by X-ray emission. Diffraction techniques do
not work well for poorly ordered phases, as these phases
yield very broad peaks that have little value for identifica-
tion purposes.
Vibrational spectroscopic techniques can be used in some
instances to identify poorly ordered phases, as these
methods detect molecular vibrations that depend on bond
energies and atomic masses rather than long-range atomic
order. The two major types of vibrational spectroscopy
are infrared absorption and Raman scattering. Infrared
absorption spectroscopy requires a long-wavelength (ca.
1-100 urn) source of light and measures the fraction of
this light that is absorbed by the sample. Water absorbs
strongly in much of this region, and so samples are best
analyzed when dry, unless internal reflectance techniques
are used. The long wavelengths used, however, are well
suited for interferometry in which the incident beam is split
in two and recombined after delaying one beam by a known
path length. The resulting interference spectrum, collected
as a function of path length difference, is then converting
using a Fourier transform to yield an absorption spectrum
in terms of wavelength. Because the entire spectrum is
collected at all times, rather than scanning through each
wavelength individually, the Fourier-transform approach
(i.e., FTIR) is the most efficient means of collecting an
infrared absorption spectrum.
Raman scattering spectroscopy measures the loss or gain
in energy of monochromatic light as a result of interactions
with molecular bonds. In theory, incident light of any wave-
length can be used to measure these changes, and so some
of the limitations of infrared absorption spectroscopy are
avoided, such as the need to study dry samples. However,
Raman scattering yields a weak signal, and fluorescence
from the sample may interfere when incident light in the
visible region is used. The recent development of Fourier-
transform Raman spectroscopy using laser light in the near
infrared region (ca 1 urn wavelength) has provided enough
incident light intensity with minimal fluorescence to make
Raman practical for many samples.
Vibrational spectroscopic techniques, while sensitive to lo-
cal bond strengths, are rarely as definitive as XRD because
the same types of bonds are found in most inorganic com-
pounds (e.g., oxide, hydroxyl, sulfate, carbonate, etc.) and
the differences among compounds are often subtle. This
approach, however, can be used to positively identify solid
solutions of poorly ordered materials, such as (Fe,Cr)(OH)3
(Amonetteand Rai,1990), based on shifts in bond energies
due to substitution. Vibrational techniques also find use for
identifying orientation and bonding of sorbed contaminants
that exist in ground water as oxyanions (e.g., arsenic; Egg-
leston et al., 1998).
Under certain circumstances, solubility can be used to
determine or at least confirm the environmentally relevant
phase controlling the availability of an inorganic contami-
nant. These circumstances require collection of equilibrium
solubility data across a range of pH (or some other primary
solubility-controlling variable), in which the concentrations
of the contaminant as well as other key species involved in
formation of aqueous complexes with the contaminant are
measured. Based on these data, and known thermodynam-
ic solubilities of the putative contaminant phase, a solubility
diagram can be constructed and compared with that for
known phases. It is critical that equilibrium be attained or
thermodynamic calculations will not be valid. Thus systems
that approach equilibrium slowly may not be suitable. Rai
et al. (1984) reviewed this approach and its applicability to
twenty-one environmentally relevant elements.
IIIB.2.4 Operationally Defined Form
A more direct approach to assessing environmental avail-
ability of a contaminant, involves determining the conditions
under which it can be dissolved and therefore mobilized
into the ground water. This empirical approach presup-
poses nothing about the chemical form of the contaminant
and thus has little predictive value in the event that ground
water conditions change. Nevertheless, it offers much in
the way of addressing economically the degree to which
a total concentration of a contaminant poses an environ-
mental risk.
The primary technique used is that of sequential selec-
tive extractions, with each successive extracting solution
offering a harsher solution environment than the one that
preceded it. The general assumption is that the earlier in
the sequence that the contaminant is released, the higher
is its environmental availability. Sequential extraction ap-
proaches are discussed in more detail in Sections IIIB.2.4.1
and IIIB.2.4.2.
The sequential (selective) extraction approach can be
improved by incorporating measurements of the rate of
release during each step in the procedure. This can be
done by successive extractions for shorter time periods
using the same reagent, by measurement of small aliquots
taken from the extraction solutions at different times during
the extractions, or by a continuous-flow extraction using a
specially designed flow cell similar to that of Wollast and
Chou (1985) used in mineral weathering studies. Despite
the apparent need, the kinetic approach has not been
widely applied to estimation of environmental availability
of contaminants, although it has been suggested for use
in assessing the availability of U in soils (Amonette et al.,
1994b).
IIIB.2.4.1 Sequential Extractions
Sequential extraction methods will perhaps be the most
practicable approaches for demonstrating contaminant
partitioning in the solid phase and for providing informa-
tion for substantiating proposed attenuation mechanisms.
These methods can be particularly useful because they
provide quantitative information on the capacity of a given
material to attenuate inorganic contaminants. They are
also advantageous because large numbers of samples can
be analyzed and compared, unlike methods that involve
spectroscopy and microscopy. However, as a cautionary
note, sequential extraction methods are not without limita-
tions. It must be acknowledged that results of sequential
extraction tests are operationally defined and users of these
tests must be aware of potential artifacts, as will be pointed
out in the following discussion.
55
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Sequential extraction procedures consist of subjecting a
given quantity of soil or aquifer sample to a series of increas-
ingly aggressive reagents under specified conditions. An
underlying assumption is that the release of components in
earlier extractions implies a higher potential for environmen-
tal mobility than components released later in the extraction
sequence. A conceptual model of the sequential extraction
approach is that solid materials consist of specific mineral
fractions that can be extracted selectively by using appro-
priate reagents and experimental conditions. As mineral
fractions are selectively dissolved, any element that they
contain will release into solution and in this way the "specia-
tion" of particulate trace metals can be determined. If it is
determined that the majority of a contaminant of concern is
present, for example, in the operationally-defined carbon-
ate fraction, then geochemical modeling efforts might be
appropriately focused on the factors that govern long-term
stability of carbonate minerals.
Perhaps the most often followed or modified chemical
extraction procedure is that of Tessier et al. (1979). Other
approaches are outlined, for example, in Dragun (1988) and
Yong et al. (1993). In the Tessier et al. (1979) method, five
distinct extraction procedures are used to recover metals
from the following sediment/soil fractions: 1) exchange-
able sites; 2) carbonate minerals; 3) metal oxyhydroxides;
4) organic matter; and, 5) the residual fraction (Table 3.1).
Metals associated with loosely bound sites (weak electro-
static attraction) are released by extraction in concentrated
salt solutions. Divalent metal chloride salts prepared at a
concentration of 1 molar are often used in this first step
(e.g., CaCI2 or MgCI2). After extraction for a set period of
time and mixing rate, the supernatant solution is analyzed
using spectroscopic methods and the concentration of an
element in solution is related back to a mass fraction of that
element associated with operationally defined exchange-
able sites. Metals associated with carbonate minerals are
removed in a second step by wet chemical extraction in a
sodium acetate-acetic acid buffer solution. This buffer (pH
5) is effective in dissolving calcite, an abundant carbonate
mineral found in natural systems. Iron and manganese
oxides are removed with hot hydroxylamine hydrochloride
solution. Organic compounds, such as humic acids and
fulvic acids, are targeted with acidic, oxidizing reagents
(hydrogen peroxide plus nitric acid; note that other published
extraction methodologies employ basic conditions to "selec-
tively" remove organic matter, for example, 3 M KOH). The
residual fraction is typically determined from a concentrated
nitric acid digestion or total analysis (by complete digestion
or by X-ray fluorescence spectroscopy) after subtracting out
the fraction of metal extracted in steps 1 through 4. The
residual fraction is considered to represent the fraction of
metals present in tightly bound matrices, for example, in
aluminosilicate matrices that are not effectively dissolved
with any of the reagents used in steps 1-4.
This extraction scheme has been adapted for testing and
validation against natural matrix standards by the Radio-
activity Group of the National Institute of Standards (MIST)
Ionizing Radiation Division (http://physics.nist.gov/Divisions/
Div846/Gp4/environ.html). This work represents an effort
to support metrology improvements in the radiochemistry
community through research and development of low-level
radionuclide Standard Reference Materials (SRMs) and
Table 3.1 Sequential extraction procedure of Tessier et al. (1979).
Step
Step 1 :
Exchangeable
Step 2:
Carbonates
Step 3:
Fe and Mn oxides/
hydroxides
Step 4:
Organic matter
Step 5:
Residual
Target Fraction
lonically bound metals and metalloids
Metals/metalloids with carbonate minerals
(e.g., calcite, dolomite)
Metals/metalloids with iron and
manganese hydroxides
Metals/metalloids bound to natural
organic matter (e.g., humic and fulvic
acids)
Aluminosilicate minerals, pyrite
Chemicals/Conditions
1 M magnesium chloride
(pH = 7.0)
1 hour, room temperature
1 M sodium acetate plus
acetic acid (pH = 5.0)
4 hours, room temperature
0.04 M hydroxylamine hydrochloride in 25%
(v/v) acetic acid (pH ~ 2)
6 hours, 96°C
0.02 M nitric acid, 18% hydrogen peroxide
(pH = 2)
2 hours, 85 °C
16 N nitric acid
2 hours, 140°C
Notes: Initial sample weight is generally 1 gram. This initial mass is used to report all metals fractions in terms of metal weight per gram of dry
sediment. Samples are generally dried (aerobically or anaerobically) to a constant mass, gently ground and homogenized in an agate mor-
tar. Samples should be constantly agitated during each extraction step. Extractions are conveniently carried out using polypropylene or Teflon
centrifuge tubes. Initial and final pH values should be measured and recorded. Supernatant solutions are collected and analyzed by atomic
absorption spectroscopy, inductively coupled plasma emission spectroscopy, or other suitable methods. Between steps, samples are centrifuged
(typically 5000 to 10000 rpm, 30 minutes) and rinsed with deionized water (5-10 mL).
56
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measurement quality assurance. The Radioactivity Group
has undertaken a systematic study to test the relative selec-
tivity of a Tessier-based extraction scheme for the speciation
of a suite of radionuclides and stable elements (Schultz et
al. 1998a; 1998b; 1998c; 1999). These tests have been
conducted with a suite of natural matrix standards (soils
and sediments) developed by NIST for the purpose of
validating radiochemical methods. While the adoption of
this extraction scheme may not be suitable for all sample
matrices (see discussion below), the development of a
uniform reference database relative to the performance
for a given extraction scheme provides a valuable starting
point for adapting an extraction scheme(s) better suited for
site-specific conditions. The availability of certified SRMs
for testing extraction schemes is critical for assessing data
quality across analytical laboratories that may be utilized
to support site characterization.
In sulfate-reducing systems, the iron sulfides mackinawite
(FeS) and pyrite (FeS2) may be potential hosts to a variety
of metals. Huerta-Diaz and Morse (1990) presented a se-
quential extraction method for the determination of metal
partitioning in iron sulfides. Their method involves the
sequential leaching of samples using 1 M HCI to recover
a "reactive" fraction (FeS; equivalent to acid volatile fraction
discussed below), 10 M HF to recover the silicate fraction,
and finally concentrated HNO3 to recover the pyrite fraction.
Moore et al. (1988) investigated the partitioning of metals
in reducing sulfidic systems using sequential extraction
techniques. They chose to use a three-step procedure:
1) 0.25 M hydroxylamine hydrochloride in 25% acetic acid
to remove metals bound to oxyhydroxides of Fe and Mn
and carbonates; 2) 0.1 M sodium pyrophosphate at pH 10
to recover the organic fraction; and 3) the sulfide fraction
was recovered using potassium chlorate plus hydrochlo-
ric acid. These fractions were compared with total metal
concentrations determined by digestion of samples in HF
plus perchloric acid.
Numerous reagents and extraction schemes have been
developed for iron minerals in soils and sediments (e.g.,
Mehra and Jackson, 1960; Chao and Zhou, 1983; Walker,
1983; Canfield, 1989; Ryan and Gschwend, 1991; Kostka
and Luther, 1994). Because of the importance of iron
minerals for sequestering metals and metalloids in the
environment, it is worthwhile to summarize the most com-
monly used wet chemical techniques for iron, and the oxide
and sulfide minerals dissolved using these various reagents
(Table 3.2). Slight variations in technique can have pro-
found influences on the outcome of chemical extractions,
for example, whether 1 M or 4 M hydrochloric acid is used
(e.g., Chao and Zhou, 1983). Consequently, in all cases
where sequential extraction methods are employed it will
be necessary to clearly document the types of reagents
used, extraction times, temperature, and laboratory pro-
cedures. Validation of laboratory procedures by spiking
Table 3.2 Summary of reagents used to selectively dissolve iron oxides and sulfides.
Extractant
Ascorbate1
HCI1-3
Hydroxylamine
hydrochloride1'4
Oxalate5
Dithionite1<2<67
Ti-citrate8
Chromous
chloride910
Composition
0.17 M sodium citrate; 0.6 M sodium
bicarbonate; 0.023 M ascorbic acid
0.5-6 M hydrochloric acid
0.04 M hydroxylamine hydrochloride in
25% (v/v) acetic acid, 96 °C
0.2 M ammonium oxalate/ 0.2 M oxalic
acid
0.11 M sodium bicarbonate/0.27 M
sodium citrate, add 0.5 grams sodium
dithionite at 80 °C, repeat dithionite
addition
0.05 M Ti(lll) citrate EDTA-bicarbonate
1 M Cr(ll) chloride in 0.5 M
hydrochloric acid; reagent prepared
using a Jones reductor
PH
8
<2
<2
2.5
5-7
7
<2
Time (h)
24
1
6
48
4
2
1
Fe minerals Extracted
ferrihydrite
ferrihydrite mackinawite
ferrihydrite mackinawite
ferrihydrite lepidocrocite
magnetite (partial) mackinawite
ferrihydrite
goethite (partial)
hematite (partial)
magnetite (partial)
ferrihydrite
goethite
hematite (partial)
pyrite
mackinawite
elemental sulfur
Notes: see references for experimental details; 1Kostka and Luther (1994); 2Canfield (1989); 3Chao and Zhou (1983); 4Tessier et al. (1979); 5Schw-
ertmann (1964); 6Mehra and Jackson (1960); 7Walker (1983); Ryan and Gschwend (1991); 9Zhabina and Volkov (1978); 1°Tack et al. (1997).
57
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extraction experiments with known quantities of minerals
and assessment of recovery will be needed in most cases
to evaluate and verify assumptions related to sequential
extraction approaches (see, for example, Chao and Zhou,
1983; Kostka and Luther, 1994; Keon et al., 2001). In this
way, the results of sequential extraction experiments can
depart from purely operational definitions and approach
the point where such results can be confidently applied to
assess mineral-contaminant associations.
Over the past 10 years, the European environmental re-
search community has developed a three-step sequential
extraction scheme, the so-called BCR method (formerly
the Community Bureau of Reference, now referred to as
The Standards, Measurements and Testing Programme of
the European Commission; see Rauret et al., 1999). The
extraction scheme has been evaluated using standard ref-
erence materials (CRB 601, sediment; CRM 483, sewage
sludge amended soil) by multiple laboratories (Ure et al.,
1993; Quevauviller et al., 1997; Rauret et al., 1999). In step
1 of the BCR scheme, metals present in ionic forms, bound
to carbonates, and in exchangeable forms are separated
(Table 3.3). In step 2, metals bound to amorphous iron
and manganese oxides are leached, while in step 3 metals
bound to organic matter and sulfides are selectively dis-
solved (Table 3.3). The use of prescribed procedures and
standard reference materials to verify performance is an
important development and a potential advantage of the
BCR extraction approach.
Assessments of environmental risk may be more suitably
focused on the water soluble and exchangeable soil frac-
tions, i.e., the most labile metal/metalloid forms released
in the first step of the sequential extraction schemes de-
scribed above. While total metals contents and pseudo
Table 3.3 BCR extraction scheme applied to 1 gram of
sample.
Step
Step 1
Step 2
StepS
Reagents/Conditions
40 ml of 0.11 M acetic
acid; 16 h at room
temperature
40 ml of 0.50 M
hydroxylamine
hydrochloride plus 0.05
M nitric acid, 16 h at
room temperature
10 ml 8. 8 M hydrogen
peroxide at room
temperature for 1 h plus
1 h at 85 °C, reduce
solution volume to near
dryness; 50 ml 1 M
ammonium acetate
adjusted to pH 2 with
nitric acid; 16 h at room
temperature
Target Fraction
Ionic,
exchangeable,
carbonates
Amorphous iron
and manganese
oxides
Organic matter,
sulfides
Notes: see Rauret et al. (1999) for detailed experimental
procedures.
metals contents (sum of all extractable forms) are valuable
in defining the extent of metal buildup in contaminated
soils and sediments, these fractions may be less useful for
assessing environmental and ecological impacts (Gupta
et al., 1996). Use of single-step extraction tests with salt
solutions of CaCI2 and NaNO3 for the determination of the
most readily bioavailable metal fraction is discussed, for
example, by Maiz et al. (1997) and Gupta et al. (1996). Al-
though single-step procedures may be useful in developing
models of ecological risk, these procedures will generally
not be adequate in developing the necessary understand-
ing of attenuation mechanisms and long-term contaminant
behavior needed to adopt monitored natural attenuation as
a site remedy.
In most cases, assessment and monitoring of natural at-
tenuation processes for inorganic contaminants will require
some type of sequential extraction procedure in Tier II. Tier
II and Tier III sequential extraction efforts might be carried
out to: i) further refine the conceptual model of natural at-
tenuation; ii) increase the spatial resolution of metal parti-
tioning data at a site; and, ill) validate extraction procedures
by spiking extraction experiments with known quantities
of reference minerals to assess method performance and
test assumptions regarding solid phase partitioning. The
selection of a sequential extraction procedure must take
into consideration site-specific factors, such as physical soil/
aquifer characteristics, redox conditions, and contaminant
type. Examples of sequential extraction approaches on a
contaminant-specific basis are presented in the element-
specific chapters in Volume 2 of this document. It is possible
that several procedures might be tried before an optimal
method is selected at any site. Samples for sequential
extraction tests should be collected near the source region
and at points moving down gradient through the contami-
nant plume and at points past the down gradient and lateral
plume fronts. Vertical resolution should cover the area most
impacted by the contaminant plume but vertical sampling
should also encompass subjacent and superjacent regions
of the aquifer. These spatial data will be needed in order
to develop a model of contaminant uptake along the flow-
path. Sequential extraction procedures should be carried
out prior to any active remediation and they may be a useful
component of annual site monitoring activities.
IIIB.2.4.2 Sequential Extraction Considerations
Many examples of selective extraction recipes can be found
in the literature. The choice of procedure will necessarily
involve consideration of project objectives and site-specific
details. Moreover, the pros and cons of using sequential
extraction methods for determining metals speciation in the
solid phase have been extensively discussed and debated
in the literature (e.g., Kheboian and Bauer, 1987; Belzile et
al., 1989; Nirel et al., 1990; Tessier and Campbell, 1991).
Sequential extractions clearly provide a very practical
methodology for getting at critical information about where
metals reside in a sample that can be linked to bioavail-
ability and geochemistry of contaminant metals. Further,
a considerable amount of work has gone into verifying the
selective extraction methods for specific elements of interest
58
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(e.g., Gruebel et al., 1988). On the other hand, a number
of complicating factors can strongly direct the outcome of
sequential extraction procedures and the interpretation of
results with regard to potential metal mobility. The two most
identifiable experimental problems with sequential extrac-
tion procedures are the non-selectivity of extractants and
potential trace element redistribution among phases during
extraction (e.g., Rendell et al., 1980; Rapin et al., 1986).
For example, metals that are released during extraction
can potentially re-adsorb to other surfaces present in the
sample at the time of extraction, or the extracting solution
itself can impact geochemical conditions that favor metal
removal by re-adsorption or precipitation. As noted previ-
ously in Section NIB.3.1, extraction results are operationally
defined and they are influenced by factors such as choice
of reagents, extraction time, and solid to solution ratios.
An example of an extraction artifact relating to arsenic is
described in Wilkin and Ford (2002). Extraction of samples
with dilute hydrochloric acid is commonly adopted to assess
metal and metalloid partitioning to the acid-volatile sulfide
(AVS) fraction of soils and sediments. This extraction test
is also used to estimate potential metal toxicity of aquatic
sediments (DiToro et al., 1990; Allen et al., 1993; USEPA,
2000b). The method takes advantage of the comparatively
high solubility of metal monosulfides like FeS, PbS, CdS,
and ZnS at low pH. However, insoluble metal sulfides of
Cu and Hg are not effectively dissolved in hydrochloric acid
(e.g., Mikac et al., 2000). Furthermore, arsenic chemistry
in su If idle systems contrasts with that of divalent metals in
that the solubility of arsenic sulfides like orpiment (As2S3)
increases with pH; solubility minima for orpiment are found
at low pH. This means that at the low pH conditions typical
of HCI extractions, precipitation of arsenic sulfide is favored.
Two potential artifacts complicate the use of low pH extrac-
tions for As solid-phase partitioning in sediments containing
acid-volatile sulfides. If orpiment is present in sediment,
it will not be efficiently dissolved with hydrochloric acid. A
more serious problem, however, relates to interpretations
of arsenic extraction data. During acid-extraction, arsenic
may be released from labile sediment components, i.e.,
loosely bound or sorbed sites. If AVS is present, arsenic
sulfide is expected to precipitate at low pH and thereby sig-
nificantly impact arsenic partitioning by transferring arsenic
from a labile and potentially bioavailable fraction to what
would be considered, using conventional interpretations, a
refractory, bio-unavailable fraction (i.e., a fraction insoluble
in hydrochloric acid).
A second example of element redistribution during chemi-
cal extraction procedures relating to lead is described in
Gerth (1990) and Ford et al. (1999). In these studies metal
partitioning to iron hydroxide was evaluated using a 2 h
extraction with ammonium oxalate. The procedure takes
advantage of the rapid dissolution of ferrihydrite in ammo-
nium oxalate relative to other more stable transformation
products of ferrihydrite, such as goethite. However, an
insoluble Pb-oxalate phase was identified as an end-product
of the chemical extraction. The formation of this insoluble
phase was clearly a consequence of the choice of reagents
used in the extraction procedure. Again without knowledge
of this artifact, interpretation of extraction results would
have likely guided the conclusion that lead was immobile
and non-bioavailable, which in fact may or may not be the
case. The magnitude of extraction non-selectivity and/or
extraction artifacts can be assessed through the addition
of specific mineral components targeted by the individual
extraction steps. These internal reference materials can
be synthesized in a manner that includes the contaminant
in a sorbed form. Synthesis methods for many of the
solid phase components targeted by the various proposed
extraction protocols are included in the cited references.
The inclusion of this form of quality control into the overall
extraction methodology will increase the level of confidence
in these analytical data.
IIIB.2.5 Attenuation Capacity
The capacity of geological materials to attenuate con-
taminants has been a subject of extensive research. The
static or batch adsorption method has often been used
to assess the capacity of soils and aquifer materials to
remove metal components from the aqueous phase. The
ease of carrying out batch-adsorption tests clearly factors
into the popularity of this testing strategy. In this method,
aqueous solutions are mixed with a given mass of solid
material for a set period of time. The aqueous solution is
next separated from the solid adsorbent material, typically
by filtration, and chemically analyzed to determine changes
in solute concentrations. The concentration of a dissolved
component before reaction with the adsorbent minus the
concentration after reaction is used to calculate the mass
of the component that has been removed by adsorption
(or some other removal process). The approach is simple
yet numerous experimental parameters impact the results
of a given batch-adsorption test (USEPA, 1992; USEPA,
1999; Jenne, 1998).
For inorganic species of concern, critical experimental
parameters include contact time, solution pH, method of
mixing, solid:solution ratio, and the concentration of other
dissolved components in the solution (e.g., Barrow and
Shaw, 1979; Roy et al., 1986). Solution composition is
of critical importance since dissolved constituents in the
background matrix of the ground water can influence con-
taminant partitioning either through competition for sorption
sites or modification of the surface charge of the aquifer
matrix. In addition, due to chemical interactions between
the aquifer matrix and the solution in which it is suspended,
it is important that solution matrix employed closely mimics
the in-situ conditions from which the aquifer matrix was
collected. For example, exposure of reduced sediments
to a solution matrix containing molecular oxygen will likely
cause significant changes in the mineralogical composition
of the aquifer matrix during the time period of the batch
test (USEPA, 2006). The results from a test performed
in this manner provide no meaningful data relative to the
assessment of in-situ sorption properties.
The methods used to prepare samples for use in laboratory-
based studies, including batch-adsorption tests, can have a
profound influence on test results. For example, oven-drying
59
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samples, although useful for homogenizing materials and
for obtaining accurate dry weights, in most cases affects
the chemical properties of sediments and may influence
the results of batch-adsorption procedures. This affect will
vary from material to material but may be the consequence
of surface area changes that come about after drying, min-
eral transformations that occur at elevated temperatures or
result from oxidation reactions. Consequently, oven-drying
is generally not an advisable practice to accelerate material
drying. Air-drying is more preferable although this method
may take a longer period of time (several days). Air-drying
can be considered to be a partial drying procedure to bring
the moisture content of a material to near equilibrium with
the atmosphere under which drying is taking place. The
duration of air-drying should be kept to the minimum time
possible and the progress should be tracked with methods
such as weighing. Anaerobic soils provide a special case.
Drying of such materials must be carried out in a anaero-
bic chamber or glove box. Drying of anaerobic materials
can be accelerated by using a stream of dry, high-purity
(oxygen-free) inert gas.
Batch-adsorption tests should be carried out under con-
stant-temperature conditions (e.g., ±3 °C) and should at-
tempt to mimic site chemical conditions. The influence of
pH on batch-adsorption tests is extremely important and
the effect of pH will vary depending on the nature of the
adsorbent and the solute of interest. Measurements of
the equilibrium pH of the soil-water suspension should be
given along with adsorption results. For anaerobic systems,
batch tests and pH measurements should be carried out
in a glove box so that air-sensitive soil components do not
oxidize and influence contaminant sorption (USEPA, 2006).
Methods for ensuring proper mixing of solid and solution
mixtures and selection of appropriate solid: solution ratios
are discussed for example in USEPA (1992).
Results of batch-adsorption tests are conveniently ana-
lyzed using linear regression tools and adsorption isotherm
equations that relate the amount of solute adsorbed to the
equilibrium concentration of the solute. Two of the more
frequently used adsorption models employ the Langmuir
Equation (Section NIB.3.2) and the Freundlich Equation
(Section NIB.3.3). The choice of one of these adsorption
models or other possible adsorption equations will typically
stem from the simplicity of the equation and from statistical
reasoning using the regression coefficient. These equations
can be used to quantitatively describe adsorption data and
be "plugged" into reaction transport models used to develop
site models (see Section ID).
Dynamic, continuous flow column experiments are a more
detailed and desirable method for obtaining metal uptake
and desorption potential. In column experiments, aquifer
materials are packed into a column apparatus, typically a
glass chromatography column equipped with Teflon end-
plate assemblies. Columns typically contain sampling
ports at the influent and effluent ends and preferably
along the length of the column. The sampling ports are
designed to allow for water sampling along the center axis
of the column. Representative solutions are then pumped
through the column using for example a high-performance
liquid chromatography pump at a flow rate selected to
approximate an average seepage velocity expected in
the field. Effluent solutions and sampling ports along the
column are then monitored with respect to contaminant
concentrations, geochemical parameters (e.g., pH, redox
potential), and volume of solution eluted from the column.
These data allow for the construction of contaminant
breakthrough curves based on the reduced contaminant
concentrations (Ceffluen/Cinput) and sample pore volume. The
results of column tests can be augmented by the application
of mineralogical characterization techniques to allow for the
identification of reaction products that can lead to insight
regarding uptake mechanisms and provide an improved
basis for predicting long-term trends.
The results of static-batch and dynamic-column tests ulti-
mately will be fed into a mathematical model (i.e., a surface
complexation model, or Freundlich adsorption isotherm) to
develop a quantitative description of contaminant sorption
to and desorption from aquifer materials. Examples of this
methodology are presented for example in Dunnivant et
al. (1992) and Kent et al. (1995), studies that explore the
transport of cadmium and chromium/selenium, respectively,
in aquifer systems.
In context of using laboratory- or field-derived data as input
to computer modeling codes, Bethke and Brady (2000)
have recently pointed out potential problems with the "Kd"
or single parameter distribution coefficient approach. When
sorption is suspected as being the dominant attenuation
mechanism, these authors argue that the simple distribu-
tion coefficient will in many cases not adequately describe
contaminant movement through aquifer systems, especially
for the ionic species typical of inorganic contaminants of
concern in this document. The distribution coefficient
simply gives the ratio of a metal ion's sorbed concentration
(mol/g sediment) to its dissolved concentration (mol/cm3).
Alternatively, Freundlich or Langmuir adsorption isotherms
or surface complexation models are available to more
accurately model field observations. The advantage of
the isotherm approach is that adsorption trends can be
tied specifically to materials collected from a specific site,
yet the method is empirical and it must be acknowledged
that extrapolation of results can not be made outside of
measured ranges. The surface complexation model does
indeed provide a more realistic description of ion adsorp-
tion from fundamental principles, yet available databases
for surface complexation constants are limited. Almost all
applications of surface complexation modeling efforts use
hydrous ferric oxide as the dominant sorbing material, this
may not be appropriate for all sites.
IIIB.3 Model Representations to Interpret
Contaminant Sorption Observations
While thermodynamic equilibrium-based geochemical
models are useful for providing boundary conditions for
estimating the extent of contaminant partitioning to aquifer
solid phases, the kinetics and reversibility of the sorption
60
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process are factors that warrant consideration on a site-
specific basis if monitored natural attenuation is proposed
as a remedial strategy for inorganic contaminants. Another
important consideration is that a purely thermodynamic
treatment of partitioning is a purely macroscopic description
and therefore, is not dependent on an atomic or a molecular
scale model which may be employed to represent sorption
reactions occurring within the aquifer (Sposito, 1981).
There are two general approaches to modeling contami-
nant sorption behavior, namely empirical and mechanistic
models. Empirical models provide a mathematical descrip-
tion of observed experimental data without necessarily
invoking a theoretical basis or microscopic model for the
observed relationship. Mechanistic models seek to de-
scribe a system based on thermodynamic principles. In
theory, the mechanistic approach is desirable because it
is more robust and widely applicable in that the effects of
changes in ionic strength and pH within a system are fully
integrated into the models. On the other hand, mechanistic
models require a rather complete chemical and physical
description of the system to be modeled and thus are still
less commonly used than the empirical models. In general,
mathematical descriptions of contaminant sorption to solid
surfaces capture the electrostatic and/or chemical forces
(or some combination thereof) that result in a net attraction
of dissolved contaminants to the solid surface. Empirical
models are derived to describe sorption trends irrespective
of the specific mechanisms, e.g., electrostatic/chemical
forces, involved in the partitioning process. In contrast,
mechanistic models attempt to describe the relationship
between properties of the solid surface and the net attrac-
tion/repulsion of the dissolved contaminant.
IIIB.3.1 Distribution Coefficient/Partition
Coefficient, Kd
The partition coefficient, Kd, is the simplest model for pre-
dicting sorption in soil systems. It is defined as the ratio of
the quantity of adsorbate sorbed per unit mass of solid to
the quantity of adsorbate in solution at equilibrium:
Kd = q/C
where q = concentration of adsorbate on the solid at
equilibrium (ug/g) and C = total dissolved concentration
remaining in solution (ug/ml).
This approach assumes that the system is reversible and
that sorption is independent of the adsorbate concentration
in the aqueous phase. Like all the empirical models the
constant Kd value does not account for changing physical
and chemical conditions in the soil. An extension of the
constant Kd is the parametric Kd model where the depen-
dence of Kdof a particular contaminant on various physical
and chemical properties of the system is determined by
stepwise linear regression analysis and polynomial expres-
sions are developed that express Kd as a function of the
relevant soil and aqueous conditions. The parametric Kd
approach is preferable in that it allows for the Kd value to
vary dependant on prevalent conditions, however, it argu-
ably involves as complete an analytical characterization of
the system as the mechanistic models while still remaining
an empirical approach.
The basic tenant of the constant Kd approach, i.e., that
partitioning is a linear function of concentration, has been
shown to be invalid in many instances. Adsorption isotherm
models have been used by soil scientists to model situa-
tions were adsorption deviates from linearity. The two most
commonly used models are the Langmuir and Freundlich
adsorption isotherm models.
IIIB.3.2 The Langmuir Model
The Langmuir model was originally developed to describe
the adsorption of gas molecules on a homogenous solid
surface. The assumptions underlying the model are that
every adsorption site is equivalent, and that the ability of
a site to bind the adsorbate is independent of whether
neighboring sites are occupied. The Langmuir equation
is expressed as:
q =
bKC
l+KC
Where q is the concentration of adsorbate on the solid and
C is the concentration in solution, b is the maximum number
of available sites for adsorption (assumes monolayer cover-
age) and K"is a constant related to binding strength. The
Langmuir equation can be rearranged to a linear form:
Let Kd = q/C
then Kd = bK - Kq
and a plot of Kd versus q should be linear with a slope of
-K and intercept of b.
Sposito (1984) reports that it is not uncommon for the
relationship between Kd and q to be convex to the q axis
rather than linear. This type of isotherm has been fit with
a two site Langmuir isotherm:
b.K.C t b2K2C
+ 1 + /CC
q=
The adherence of an adsorption isotherm to a two site
Langmuir model has been interpreted as evidence for
two discrete binding sites on the solid phase, however, no
mechanistic interpretation can really be inferred from the
goodness of fit of sorption data to these isotherms.
IIIB.3.3 The Freundlich Isotherm
The Freundlich isotherm has the form:
q = ACf
Where A and (3 are adjustable parameters and (3 can have
a value between 0 and 1. Hence a plot of log q vs. log C
should be a straight line with an intercept of log A and slope
of (3. Note that when (3 = 1, then the Freundlich equation
reduces to the linear Kd.
Sorption isotherms have been widely used to describe and
predict adsorption of a contaminant in soil and sediment
systems, however, as stated previously they are purely
61
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empirically derived relationships and the validity of the cal-
culated parameters can only be expected to hold within the
bounds of the experimental data used to create the initial
isotherm. They also provide no mechanistic interpretation
of the reactions governing the solubility of a contaminant
in a system; they are insensitive to the mode of sorption
(i.e. precipitation or adsorption). In view of these facts
mechanistic models of the soil sediment system have been
developed. The mechanistic approach uses thermodynamic
relationships to model aqueous speciation and electric
double layer theory to model changes in surface charge
as a function of pH, electrolyte concentration and valence.
Mechanistic models are potentially much more robust than
adsorption isotherm approaches and, once the model has
been established, potentially much more powerful in their
predictive capabilities with respect to changing physical
and chemical conditions within the system.
IIIB.3.4 Mechanistic Models for Predicting
Sorption - Surface Complexation
A variety of mechanistic approaches have been applied to
provide a molecular description of adsorption in soil-water
systems. A number of approaches are based on a math-
ematical description of the distribution of ions in the vicinity
of a charged surface, and the surface charge and poten-
tial in the interfacial region. These mathematical models,
termed surface complexation models (SCMs), capture the
influence of electrostatic forces between the solid surface
and charged ions within solution as well as the influence
of chemical interactions between these two entities that
leads to the formation of coordinative bonds. The various
models that are commonly employed to describe surface
complexation reactions differ in their description of the
distribution of charge at the solid-water interface and how
ions are distributed in the aqueous layer that bathes the
solid surface. Details of the various models are available
from several sources and should be consulted by those
unfamiliar with the implementation of surface complexation
modeling to describe contaminant solid-solution speciation
(e.g., Sposito, 1984; Davis and Kent, 1990; Stumm, 1992;
Goldberg, 1995). A brief synopsis of the various models
is provided in Table 3.4. In general, two different types of
surface complexes are proposed in the various models: 1)
inner-sphere surface complexes in which the contaminant
forms a bond with a surface functional group, and 2) outer-
sphere surface complexes where the contaminant partitions
to the surface via electrostatic attraction (similar to the
process of ion exchange).
A key difference between SCMs and empirical models is the
employment of thermodynamic concepts to model chemi-
cal reactions occurring at the solid surface. By proposing
specific chemical reactions for the partitioning of solution
ions to the solid surface it becomes feasible to account for
the influence of the bulk solution composition such as pH,
which can exert an influence on both surface charge as well
as solute speciation. In addition, the bulk water composi-
tion can influence contaminant sorption via the presence
of ions in solution that compete for available surface sites.
Examples of surface complexation reactions that may be
Table 3.4 Synopsis of the various surface complex-
ation models (SCMs) commonly employed to
describe solute partitioning to solid surfaces.
Surface
Complexation
Model
Constant
Capacitance
Model
(CCM)
Diffuse Layer
(DIM)
Triple Layer
Model
(TLM)
Nonelectrostatic
Model (NEM)
Description
of Solute
Partitioning
to Solid
Surface
Inner-sphere
complexes
Inner-sphere
complexes
Inner- and
outer-sphere
complexes
Inner-sphere
complexes
Influence of
Surface Charge
Exponential
electrostatic term
that modifies
value of surface
complexation
constant
Exponential
electrostatic term
that modifies
value of surface
complexation
constant
Two exponential
electrostatic terms
that modify values
of inner-sphere
and outer-
sphere surface
complexation
constants
postulated within an SCM are as follows:
=SH + Zn2+ <-» =S-Zn+ + H+
(=SH represents functional group on FeS)
=SH + Mg2+ o =S-Mg+ + H+
=SH + OH- <-» =S- + H2O
=FeOH + Pb2+ o =FeOPb+ + H+
(=FeOH represents functional group on FeOOH)
=FeOH + Ca2+ <-» =FeOPb+ + H+
=FeOH + H+
=FeOH
A mass action equation can then be written for each of
these reactions (similar to solution complexation reactions)
and solved to calculate conditional surface complexation
constants. For SCMs that account for the influence of
electrostatic interactions, the conditional surface complex-
ation constant is typically modified via multiplication by
an exponential electrostatic factor in order to calculate an
instrinsic surf ace complexation constant.
62
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The reactions describing ion partitioning to a solid surface
provide the physicochemical context for assessing changes
in sorption that may accompany changes in ground-water
chemistry. Several observations are evident from the ex-
ample reactions shown above:
1) pH can influence the speciation of surface func-
tional groups as well as the surface charge,
2) ion sorption can influence surface charge, and
3) major ions in solution (e.g., Mg2+ and Ca2+) can
compete with the contaminants of concern (e.g.,
Zn2+ and Pb2+) for available sorption sites.
Thus, an SCM provides one with the ability to project how
changes in solution chemistry can impact contaminant
uptake; a significant limitation of empirical partitioning
relationships. While this flexibility provides a powerful tool
for examining the evolution of a contaminant plume, current
implementations of most SCMs is limited by the heteroge-
neous nature of solid surfaces within an aquifer (types of
sorption sites) and the ability to predict surface charging
behavior. Recent efforts have been employed to develop
nonelectrostatic surface complexation constants for site-
specific descriptions of subsurface contaminant transport
(e.g., Kohler et al., 2004). This approach will most likely be
successful for contaminants that predominantly form strong
chemical bonds with available solid surfaces.
///B.3.5 Mineral Solubility
Characterization of the solid phase will, in most cases, be
an integral component of monitored natural attenuation
assessment and application. Some insight regarding the
mineralogical composition of aquifer systems can be ob-
tained by analyzing solution compositions. Investigations
of mineral solubility in aquifer systems usually concentrate
on the question: does a mineral control the concentration of
a particular element, and, if so, what is the identity of the
mineral? In order to answer these questions, geochemical
models are frequently employed to calculate ion activities
and mineral saturation indices. In aquifer systems, some
chemical reactions are sufficiently fast that equilibrium
relationships are immediately attained. For example, pro-
tonation/deprotonation reactions of acids and bases and
ion pairing reactions are fast chemical reactions. Other
reactions, in particular those reactions involving solids,
can proceed so slowly that equilibrium is not attained even
after decadal time periods. Yet, equilibrium is a practical
reference point and equilibrium relationships are useful for
predicting reactions that are likely or unlikely to occur.
Mineral solubility is influenced by the ionic strength of
solution. Unless conditional equilibrium constants are em-
ployed, solution activity models are a necessary component
of geochemical modeling efforts. Activity models account
for the non-ideal behavior of solute ions in aqueous solu-
tions. Non-ideal behavior is a consequence of electrostatic
interactions between water molecules and charged solute
ions. Methods for computing individual ion activity coef-
ficients are presented in numerous textbooks (e.g., Stumm
and Morgan, 1981; Nordstrom and Munoz, 1986). For
most ground water studies of low ionic strength waters, the
extended Debye-Huckel equation or the Davies equation
will provide reasonable activity models. In these equations,
activity coefficients are estimated based on input values for
solution ionic strength and temperature. For concentrated
waters that are high in total dissolved solids, virial methods
such as the Pitzer model are available for estimating ion
activity coefficients (Langmuir, 1997).
The unit of concentration for dissolved species most fre-
quently used for aqueous solutions is molality, m. (mol/kg).
Analytical concentrations are often expressed in mass
based units, e.g., ppm (parts per million). A useful con-
version is:
Cone, in ppm = Cone, in mol/kg x
formula weight in g/mol x 1000
The effective concentration or activity of a dissolved spe-
cies, a., is given by:
Where y is the ion-specific activity coefficient and (y° m°)
refers to the standard state, which for aqueous solutions
is typically chosen as an ideal, 1 molal solution, i.e., both
y° and m° are equal to 1 .
In order to evaluate whether a ground water is oversatu-
rated, undersaturated, or at equilibrium with a particular
phase, geochemical speciation models are of practical
use. As an example, consider the solubility expression for
cadmium sulfide (greenockite):
CdS(s) + H+ = Cd2+ + HS-
The mass-action expression that applies to the equilibrium
is:
a a
I/ _ Cd2+ HS~ _ i Q-14-4
Ground water may or may not be at saturation with respect
to greenockite, depending on whether the phase is indeed
present, available surface area, residence time of water,
and kinetic factors that may impede dissolution and/or
precipitation. If the water is at equilibrium, then the ion
activity product, Q, should be the same as the equilibrium
constant, i.e.,
where the activity of CdS is taken to be unity. Calculation
of the saturation index (SI) for a water can then be used to
determine if the solution is undersaturated, at equilibrium,
or oversaturated with respect to precipitation (see Section
IIB.2.1, Table 2.3).
It is important to point out that solubility products for pre-
cipitates, such as cadmium sulfide, depend on whether
the solid is freshly precipitated (typically disordered) or
crystalline (well ordered). At 25 °C, Daskalakis and Helz
(1992) report a solubility product of 10'1436 for crystalline
cadmium sulfide (greenockite). Wang and Tessier (1999)
63
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give a value of a freshly prepared CdS precipitate of 10'14-15.
Hence, freshly precipitated cadmium sulfide is metastable
with respect to greenockite since:
CdS(disordered) = CdS(greenockite) Kr = 10021
and
AG° = -aflTlog K= -1.20 kJ/mol
If it is possible to calculate activities of free ions from a given
water analysis, then it is a relatively simple matter to cal-
culate the degree saturation of a large number of possible
mineral phases, keeping in mind that temperature and ionic
strength are needed to correct the ion activity product.
In the example above, pH and the molal concentrations
and ion activity coefficients for Cd2+ and HS" are needed
in order to compute the ion activity product. If complexes
of cadmium are present, then mcd2+ is not the same as the
total concentration of dissolved cadmium. If we consider
only mononuclear bisulfide complexes,
mcd,total = mr
m
, + m „
CdHS+ Cd(HS)"
+ m
Cd(HS)~
Depending on the solution composition, cadmium com-
plexes with chloride, bicarbonate, hydroxide, orsulfate may
be important in addition to bisulfide complexes. Similarly,
mHS- is not the same as the total concentration or analytical
concentration of dissolved sulfide. Both protonated and
deprotonated forms of sulfide and complexed forms of
sulfide may be present. Speciation calculations, therefore,
are critically dependent on the completeness and quality of
thermodynamic constants in the thermodynamic database
used for geochemical modeling (see Section ID).
In cases where mineral solubility is controlled in part by
redox conditions, calculation of the saturation index is not
straightforward. For example, when an aqueous solution
has attained saturation with respect to a mineral such as
the iron oxide magnetite (Fe3O4), then the reaction
Fe3O4 + 8H+ = Fe2+ + 2Fe3+
is at reversible equilibrium when
4H2O
a 7+a
Fe2+ Fe
- = io42=/c
H+
Magnetite solubility depends therefore on pH and redox. It
follows that to calculate the saturation index for magnetite,
activities of both the ferrous and ferric ions are needed input
variables. Accurate measurement of ferric iron concentra-
tion at circumneutral pH is a notoriously difficult task. There
are several ways to approach this problem:
• Measure separately both Fe(ll) and Fe(lll) in a water
sample and estimate activity coefficients in order to
calculate the activities of Fe2+ and Fe3+
• Use Eh as a master redox variable to fix Fe(ll):Fe(lll)
ratios
• Measure Fe(ll), calculate Fe(lll) by assuming control
by the solubility of Fe2O3, FeOOH, or Fe(OH)3
• Use some other measured redox pair to fix Fe(ll)/
Fe(lll)
There are methods available to determine separately the
concentration of ferrous iron and ferric iron (e.g., To et al.,
1 999); however, these methods are most effective for use in
low pH waters, such as aquatic systems impacted by mine
wastes. At near-neutral pH, the concentration of Fe(lll) is
typically very low so that analytical detection limits constrain
the use of direct measurements.
The second method is to use the measured Eh of a water
to estimate the Fe(ll)/Fe(lll) ratio in solution. Applying the
Nernst equation to the iron(ll/lll) couple, we have
RT
/' =0.770 -0.0592 log
Using this approach, Fe(ll)/Fe(lll) can be calculated from
a potential measurement using a platinum electrode. Nor-
dstrom et al. (1979) conclude that this approach is most
effective at pH <6 and at total iron concentrations exceeding
10"6 molal. A third method is to assume that the solubil-
ity of a ferric-bearing mineral controls the activity of Fe3+.
For example, consider the solubility expression of hydrous
ferric oxide:
Fe(OH)3 + 3H+ = Fe3+ + 3H2O K = 10'36 6
(Langmuir, 1996)
So that,
log ape3+ = log K - 3 log aH,,0 - 3pH = -36.6 - 3pH
In this example, the activity of Fe3+ can be estimated based
on a measurement of pH. In the last example, a separate
redox pair, e.g., As(lll)/As(V), can be used to fix the Fe(ll)/
Fe(lll) ratio in solution by applying the Nernst equation. The
correctness of any one of these approaches relies on how
closely the assumptions are obeyed for a given system.
IIIB.3.5.1 Coprecipitation Reactions
The concept of a solid solution implies an isomorphic sub-
stitution, for example, regular substitution of Mn for Ca in
a carbonate mineral. Minerals formed in the environment
often contain substitutional impurities and coprecipitation
is likely to be a primary natural attenuation mechanism.
Thermodynamic models of solid solutions indicate that the
solubility of a component becomes greatly reduced as that
component becomes a constituent in a solid phase. In other
words, solid solutions in binary systems, for example, are
less soluble (more stable) than pure end-member composi-
tions. A considerable effort has gone into understanding the
thermodynamic and kinetic factors that control the formation
of solid solutions from aqueous solutions (e.g., Lippmann,
1977; Busenberg and Plummer, 1989; Glynn et al., 1990;
Glynn and Reardon, 1990).
Laboratory and field studies indicate that cadmium is ef-
fectively removed by calcerous materials (e.g., Dudley et al.,
1988; Davis et al., 1987). The solid solution (Ca,Cd)CO3 is
complete and the endmember solubility products differ by
three orders of magnitude at room temperature:
CaCO, = Ca2+ + CO 2- log Kr m = -8.3
3 3 " OaOOo
and
64
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CdC03 = CcP + CO/- log K-CdC03 =-11.3
Combining these solubility expressions we obtain an ex-
change reaction:
CaCO, + Cd2+ = CdCO, + Ca2+
o o
The equilibrium constant for this type of reaction is often
referred to as a distribution constant, D, and is given by the
quotient of the solubility products of CaCO3 and CdCO3:
D =
= io3
The activity ratio of the solids by definition can be replaced
by the ratio of the mole fractions multiplied by activity coef-
ficients:
D =
Therefore, the amount of substitution of Cd into calcium
carbonate is a function of the solubility product ratio of
CaCO3 to CdCO3, the solution activity ratio of Cd2+ to Ca2+,
and a term that represents activity coefficients of the solid
solution components:
(/
_ "CdCOj
3Cd2+ J YcdC03 I
3ca2+J\TcdC03J
In general practice the first two terms, solubility product
ratio and the solution activity ratio, are straightforward to
obtain. Activity coefficients in the solid phase in most cases
differ markedly from 1. In principle, the activity term can
be determined experimentally by measuring the mole frac-
tion ratio over a range of solution activity ratios at constant
temperature.
An analysis of solution concentration data alone will
generally not be adequate to confirm any precipitation or
coprecipitation mechanism of attenuation. Multiple lines of
evidence, including solution and solid-phase data, which
supports a specific natural attenuation mechanism will in
most cases be necessary to pursue monitored natural at-
tenuation as a partial or sole cleanup remedy.
IIIB.3.5.2 Thermodynamic Data
Site characterization and remedial investigations that
involve contaminant transport in the environment are usu-
ally accomplished, in part, with geochemical modeling.
Consequently, thermodynamic properties are essential
components to geochemical modeling efforts. Normally, the
thermodynamic constants used in modeling exercises are
taken from published compilations of such properties (e.g.,
Wagman et al., 1982). In some cases these compilations
contain minimal documentation as to the source or qual-
ity of compiled data. In other cases, compilations contain
outdated information. Users of thermodynamic databases
can easily be misled into believing that modern databases
are up to date in terms of data quality and completeness.
As an example to illustrate problems noted above, a spe-
cific example of orpiment (As2S3(s)) is discussed. Eary
(1992) listed a number of values of the Gibbs free energy
of formation, AGf°, of orpiment, each from a separate lit-
erature source. The values tabulated by Eary (1 992) range
from -168.8 kJ/mol to -90.7 kJ/mol. This wide range in
Gibbs free energy values equates to equilibrium constants
describing orpiment dissolution and precipitation that can
span almost 7 orders of magnitude. For example, consider
the solubility expression for orpiment to form arsenite in
solution:
0.5As2S3
3H2O = As(OH)3° + 1 .5HS" + H+
An equilibrium constant for this reaction can be computed
from the AGf° values for orpiment given above and fixed
values of AGf° for H2O(I), As(OH)3°, and HS" of -639.8 kJ/
mol, -237.18 kJ/mol, and 12.05 kJ/mol, respectively. The
more negative AGf° value for orpiment gives a log K value
for the above reaction of -30.5, while the greater value gives
a log K" value of -23.7. The pH dependent trend of orpi-
ment solubility resulting from these differing log K" values is
shown in Figure 3.5. The solubility diagram is constructed at
ZH2S concentration of 1 0~5 5 m, so that thioarsenite species
have a negligible contribution to orpiment solubility (Wilkin
et al., 2003). It is clear that depending on the log K value
employed in reaction modeling, solutions will have vastly
different saturation indices and one would reach different
conclusions about possible attenuation mechanisms for
arsenic.
-2-
-4-
i *A
§> -8-1
-10-
-12-
-14
As2S3(s) solubility based
on AG, = -90.7 kJ/mol
As2S3(s) solubility based
on AG. = -168.8 kJ/mol
10
12
pH
Figure 3.5
pH-dependent solubility trend of orpiment
predicted using two different Gibbs free
energy of formation values (see text). The
model curves correspond to a constant
Z/-/2S concentration of 10'ss m. The data
points are measured solubility data taken
from Webster (1990). The experimental
data are consistent with a A.G° for orpi-
ment of approximately -90 kJ/mol, yet most
thermodynamic compilations adopt a value
of-168.8 kJ/mol. Application of the lower
value will vastly over predict the stability of
orpiment.
65
-------
Which value of AGf° for orpiment is more appropriate?
The most recent solubility and calorimetric studies seem
to support a AGf° formation value close to -90 kJ/mol for
crystalline orpiment (Webster, 1990; Eary, 1992; Johnson
et al., 1980). The key point is that users of geochemical
models should be familiar with the sources of thermody-
namic data used and attempt to evaluate whether constants
in thermodynamic compilations are in reasonable agree-
ment with recent experimental evidence or whether there
may be discrepancies that could impact how model results
are interpreted.
IMC. Characterization of System Redox and
Underlying Microbial Processes
Oxidation-reduction processes affect the chemical compo-
sition of ground water and impact the aqueous and solid
phase chemical speciation of inorganic contaminants. Thus,
characterizing the redox status of ground water systems
will likely play an important role in understanding controls
on contaminant attenuation. As stated previously in Sec-
tion IIC, the subsurface microbiology within a plume will,
in part, influence the predominant redox characteristics of
the system. Microbial processes may influence the redox
chemistry of important components within the aquifer,
such as iron and sulfur, participating as reactants within
attenuation reactions that result in contaminant immobili-
zation or degradation. Thus, it is recommended that site
characterization include both assessment of the prevailing
subsurface redox chemistry impacting contaminant trans-
port, as well as more explicit analysis of specific microbial
processes for sites at which the plume conditions support
microbial activity that differs from ambient conditions. For
example, a more detailed microbial community analysis
may be warranted in order to improve the reliability of as-
sessing attenuation capacity for sites where contaminant
source characteristics govern aquifer geochemistry within
the down gradient plume. The following discussion provides
approaches and tools available to characterize the overall
redox status of site ground water, as well as the specific
microbial communities that support the observed redox
chemistry within the plume.
IIIC. 1 Process Iden tification
Identification of a redox-mediated attenuation process starts
with knowledge of the aqueous and solid-phase species
present. Although analysis of the contaminant concentra-
tion in ground water at various distances from the source
can indicate attenuation, inclusion of data for pH, EH, major
solutes, and redox-sensitive species such as Fe2+, O2, and
H2S is needed to demonstrate that redox processes may
be involved.
Further evidence for redox-mediated processes may be
derived from determination of the mineral species present.
The absence of Fe and Mn oxides and the presence of
Fe sulfides, for example, suggest that the aquifer is pre-
dominantly reducing, whereas the inverse of this situation
suggests an oxidizing aquifer. A variety of techniques can
be used for mineral identification and is detailed elsewhere
in this document (Section NIB.2). A major consideration
related to mineralogical characterization relates to the
preservation of original mineralogy in aquifer materials by
preventing contact with oxygen during sample collection and
removal of pore water. For example, iron-bearing minerals
may exist in a reduced state within the saturated zone. Yet
such minerals that are stable in reducing environments
are subject to significant alteration upon exposure to oxy-
gen. Solid phase structural and chemical transformations
are commonly mediated or facilitated by the pore water.
Thus, removal of pore water may act to retard or impede
transformation. In order to preserve redox characteristics
of samples collected in the field, cores materials should
be immediately capped and frozen. Sample freezing can
be accomplished either by submersing in liquid nitrogen or
placement in a portable freezer located in the field. Follow-
ing transport to a laboratory setting, frozen materials should
be thawed under an oxygen-free or inert atmosphere,
e.g., within an anaerobic glove box. Extended periods of
storage should be avoided, since sample mineralogy will
alter with time during approach to a new equilibrium state.
Procedures for the collection and processing of aquifer
solids to minimize alterations to mineralogy are outlined
in several EPA documents (USEPA, 2002; USEPA, 2005b;
USEPA, 2006).
In some situations, intense local microbial activity may be
entirely responsible for the redox status of the aquifer. The
nature of the active microbial population (e.g., iron-reducing,
sulfate-reducing, or sulfur-oxidizing bacteria) can often be
inferred from geochemical data. Thus, trends in the concen-
tration of organic substrates (soluble organic C) and their
metabolites (e.g., H2, H2S, CH4, CO2, NO2', HS', Fe2+) can
indicate whether and which microorganisms are active in a
particular subsurface region. In some instances, direct and
specific determination of microbial population by culturing or
genetic analysis (e.g., messenger ribonucleic acid profiles)
of aquifer solids extracts may be warranted.
Although the best support for a particular redox process
comes from direct examination of reactants and products
at the site, a strong case for the process can be made with
laboratory tests using core materials taken from the site.
These tests involve batch or column studies in which the
geochemical inputs and conditions are comparable to those
at the actual site, and evidence for the process is obtained
from analysis of the solution phase (supernate or effluent)
during the test and of the solid phases at its conclusion.
Attenuation obviously must be demonstrated through the
solution-phase data, and direct identification of the attenu-
ated form of the contaminant provided from the solid-phase
data. As described below, this same general approach with
suitable modifications can be used to determine capacity,
kinetics, and stability.
IIIC. 1.1 Redox Measurements
Measurement of redox parameters in ground water is inher-
ently challenging due to the fact that a steep redox gradient
is often present between the sampling location (subsurface)
and the location where the particular measurement is made
66
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(surface). Consequently, special care must be taken to
preserve the redox integrity of ground-water samples. In
some cases water samples can be preserved by using inert
gases or by acidification because oxidation rates generally
decrease substantially with decreasing pH. In other cases
where sample preservation is not possible or practical, it is
recommended that analyses be carried out in the field at the
time of sample collection (e.g., for dissolved oxygen and Eh).
Summaries of methods used to determine redox status can
be found in, for example, USEPA (2002), Baedecker and
Cozzarelli (1992), and Walton-Day et al. (1990). Table 3.5
provides a list of parameters that can be used to assess
the redox status of ground water systems.
Other laboratory approaches include selective dissolution
of reactive fractions (see Section NIB.2.4.1). For example,
reactive Fe2+ (i.e., that which is sorbed as well as that pres-
ent in sulfides carbonates, hydroxides, and green rusts)
can be estimated by extraction with 0.5 M HCI followed by
complexation with ferrozine (Roden and Zachara, 1996;
Amonette et al., 2000). From the oxidizing side, Mn(lll,IV)
oxides and poorly ordered Fe(lll) (hydr)oxides can be de-
termined by extraction with hydroxylamine solutions (Chao,
1972; Chao and Zhou, 1983; Ross, 1985).
The direct reactions and laboratory tests, however, measure
only the redox-buffering capacity at a single point in time.
Microbial activity, through conversion of organic C, can
create and replenish the reductive capacity of a site. Thus,
more accurate measurements of reductive capacity may
necessitate consideration of inputs of dissolved constituents
from source areas or up gradient portions of the aquifer
such as dissolved organic C, as well as terminal electron
acceptors such as oxygen, nitrate, and sulfate, in order to
assess ability of the microbial community to maintain or
degrade the redox capacity of a site.
111C.2 Capacity
Once the operative redox-mediated attenuation process has
been identified, an assessment of its capacity to attenuate
the contaminant is needed. The primary soluble electron
donor in most ground waters is dissolved organic carbon,
although, under some circumstances, sulfide (H2S, and HS")
Table 3.5 Ground-water redox parameters and measurement approaches.
Parameter
Oxidation-reduction
potential (ORP)
Dissolved oxygen
Dissolved hydrogen
Iron speciation
Sulfur speciation
Nitrogen speciation
Arsenic speciation
Chromium speciation
Selenium speciation
Measurement Approach
Combination platinum electrode with Ag/AgCI reference electrode; KCI filling solution.
Electrode performance is determined using reference solutions (e.g., Zobell's solution,
hydroquinone).
Membrane-covered electrodes; colorimetric tests, modified Winkler titration. Electrode
performance is determined using air-saturated water and sodium sulfite solutions.
Sample collection in glass vessel and analysis by gas chromatography/reduced gas
analyzer.
Ferrous iron determined by colorimetric analysis (e.g., ferrozine, 1,10-phenanthroline).
Ferric iron determined by adding reducing agent and measuring ferrous iron, and/or
by determination of total iron and subtracting Fe(ll). Measurement made in the field or
preservation required.
Sulfate and other sulfoxyanions (sulfite, thiosulfate) are typically determined by ion
chromatography or capillary electrophoresis. Sulfide can be determined by colorimetric,
gravimetric, coulometric, or voltametric methods. Measurement made in the field or
preservation required.
Colorimetric, chromatographic, and potentiometric methods available for nitrate, nitrite,
and ammonium. Preservation of sample is recommended.
Anion exchange, chromatographic, and hydride generation methods are available for ar-
senite, arsenate, and several organic forms of arsenic. Preservation is recommended.
Colorimetric, exchange, and voltametric methods are available for determination of Cr(VI)
and Cr(lll).
Selenium oxyanions can be determined by ion chromatography or capillary
electrophoresis. Preservation of sample is recommended.
67
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and Fe2+ species can dominate. Important electron accep-
tors include oxygen, nitrate, nitrite, and sulfate. Thus, in
addition to the contaminant, ground-water analyses should
include measurements of these constituents. Equally
important is a hydrological assessment of ground water
flow rates, which, when combined with the concentration
data, allows an estimate of the average influx of oxidants
or reductants to the site. In addition, quantification of the
accessible redox-buffering capacity of the aquifer solids
may be important, since these solids may represent a
significant fraction of the capacity of the aquifer. Meth-
ods for characterizing the oxidation capacity and reducing
capacity of aquifer solids are listed in Table 3.6. These
methods represent examples of approaches that have been
tested and documented in the literature. For site-specific
applications of these methods, some method development
or modification may be required to obtain optimal results
(USEPA, 2002).
Laboratory tests can be performed in which a known oxidant
[e.g., O2(a or Cr(VI)] or reductant [e.g., H2S(a,] is reacted
with the aquifer solids under controlled environmentally
relevant conditions (Fruchteretal., 1996; Istoketal., 1999).
The quantity of this reagent consumed by reaction is then
expressed in terms of the mass or volume of the aquifer
solids. Rough estimates of maximum reductive capacity
present in a soil can also be obtained from digestions using
acidic Cr(VI) solutions. Thus, adaptations of Cr(VI) titration
methods for organic C in soils (e.g., Nelson and Sommers,
1996) are suitable as they include the contributions of Fe(ll)
and sulfides to overall reductive capacity. Other labora-
tory approaches include selective dissolution of reactive
fractions (see Section IIIB.2.4.1) with the assignment of
a specific mass-based oxidation/reduction capacity to the
phase quantified by extraction. For example, reactive Fe2+
(i.e., that which is sorbed as well as that present in sulfides
carbonates, hydroxides, and green rusts) can be estimated
by extraction with 0.5 M HCI followed by complexation
with ferrozine (Roden and Zachara, 1996; Amonette et
al., 2000). From the oxidizing side, Mn(lll,IV) oxides and
poorly ordered Fe(lll) (hydr)oxides can be determined by
extraction with hydroxylamine solutions (Chao, 1972; Chao
and Zhou, 1983; Ross, 1985).
The direct reactions and laboratory tests, however, mea-
sure only the redox-buffering capacity at a single point in
time. Microbial activity, through conversion of organic C,
can create and replenish the reductive capacity of a site.
Thus, more accurate measurements of reductive capacity
Table 3.6 Methods that may be employed for estimating the oxidation and reduction capacity for solid materials
(from USEPA, 2002).
Method
Cr(ll) (oxidation capacity)
Digestion with Ti(lll)-EDTA
(oxidation capacity)
Titration/digestion with
dithionite solution
(oxidation capacity)
Chemical Oxygen Demand
by digestion with acid
dichromate
(reduction capacity)
Digestion in hydrogen
peroxide solution
(reduction capacity)
Dissolved oxygen
consumption in
air-saturated water
(reduction capacity)
Source
(Barcelona and Holm, 1991 a)
(Barcelona and Holm, 1991b)
(Barcelona and Holm, 1992)
(Ryan and Gschwend, 1991)
(Loeppert and Inskeep, 1996)
(Williams et al., 2000)
(USEPA, 1979)
(Barcelona and Holm, 1991 a)
(Barcelona and Holm, 1991b)
(Barcelona and Holm, 1992)
(Nelson and Sommers, 1996)
(Williams et al., 2000)
Comments
Most aggressive but oxygen-free atmosphere
recommended; high estimate
Developed for extraction of Fe oxides; applicability
limited to iron oxide dominated sediments
Valid for only specific remedial technology; targets
Fe oxides
Precipitate coatings if pH not buffered; high
estimate
Developed for quantifying organic matter content
Dynamic column test with mathematical simulation;
test design must minimize gas diffusion from
external sources; time-consuming, but realistic
68
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may necessitate consideration of inputs such as dissolved
organic C, as well as terminal electron acceptors such as
oxygen, nitrate, and sulfate, in order to assess ability of
the microbial community to maintain or degrade the redox
capacity of a site. For this analysis, it may be necessary
to evaluate microbial response to dissolved constituent
inputs through sampling of aquifer solids for the purpose
of conducting microcosm studies to examine microbial ac-
tivity and/or contaminant attenuation. Recommendations
for the design and implementation of microcosm studies
are provided in USEPA (1998). Evaluation of microcosm
response to variations in electron donor/acceptor concen-
trations in solution, relative to contaminant attenuation,
provides means for directly assessing the limits in these
reactant concentrations under which attenuation remains
viable. For situations in which contaminant degradation is
the primary attenuation process, microcosm measurements
will likely include determination of trends in contaminant
loss as well as the increase of degradation products. For
situations in which contaminant immobilization dominates,
it is recommended that microcosm characterization include
determination of the quantity and solid phase speciation
of contaminant sorption. The degree to which microcosm
studies replicate subsurface conditions within the plume
may be assessed through comparison of similarities (or
lack thereof) between measured aqueous and solid phase
chemical parameters based on measurement of aquifer
microcosm properties (e.g., water chemistry, mineralogy).
An additional line of evidence to support capacity assess-
ments includes development of a reaction or reactive trans-
port model that incorporates quantitative description of the
processes that control contaminant attenuation. Require-
ments for model construction and parameter inputs have
previously been specified in Section ID. The utility of this
type of analysis is the ability to quickly assess a range of
ground-water conditions that may influence the efficiency
of modeled attenuation reactions. However, as previously
noted, the degree of uncertainty in model predictions will be
constrained by the accuracy of parameter inputs to repre-
sent aquifer conditions within the plume. Thus, verification
of model performance is warranted to demonstrate the
ability of the model to reproduce measured ground-water
conditions prior conducted model tests of aquifer capacity
to support attenuation.
///C.3 Stability
If a redox-mediated attenuation mechanism has been identi-
fied and a reasonable estimate of the capacity of the aquifer
to attenuate the contaminant has been made, evaluation
of the stability of immobilized contaminants is needed to
assess the potential for contaminant remobilization due to
anticipated changes in ground-water chemistry. This com-
ponent of the site characterization effort may include direct
measurements of contaminant stability via laboratory- or
field-based evaluations, which could be supplemented with
implementation of reaction or reactive-transport models
that explicitly consider the solid-phase speciation of the
contaminant. The ultimate goal of this effort is to gauge the
response of the aquifer, from the perspective of contami-
nant remobilization, to changes in aquifer redox status that
may be driven by future increases in the influx of dissolved
components such as oxygen or the cessation of microbial
processes that accompany decreased influx of degradable
organic contaminants.
Contaminant stability may be estimated through labora-
tory tests constructed using aquifer solids collected from
within the zone of contaminant attenuation. Controlled
tests could be then be devised that evaluate contaminant
response to changes in specific ground-water parameters
that may result in release of the contaminant from aquifer
solids. For sites in which the ground-water chemistry differs
significantly within the plume compared to up gradient or
ambient conditions, this may entail exposure of the aquifer
solids to ambient ground-water samples. Alternative ap-
proaches may include systematic variation of one or more
parameters identified as being critical to the stability of the
form in which the contaminant is immobilized. For example,
this may involve systematic variation in ground-water
sample pH over a range that captures current conditions
as well as anticipated conditions that may be reflected by
the pH measured in background wells installed within the
aquifer. Other parameters that might be assessed include
those that reflect conditions that might develop as a result
of potential land-use changes, including influxes of dis-
solved constituents that might compete for adsorption sites
and/or may form soluble complexes with the contaminant.
An alternative approach to assessing the stability of an
immobilized contaminant may include the implementation
of in-situ studies using single-well push-pull tests (Istok et
al.,1997; Haggerty et al., 1998; Senko et al., 2002). These
tests can similarly be devised to evaluate contaminant
response to changes in ground-water chemistry through
manipulation of specific parameters via mixing of synthetic
solutions with ground water retrieved from the well and
subsequently re-injected into the aquifer. The re-injected
water is then allowed to react with the aquifer solids for
short periods of time (days to weeks), and then several pore
volumes are withdrawn and analyzed to assess the fate of
the contaminant and/or analyze potential by-products that
result during a re-mobilization reaction. Comparison of
reagent concentrations with those of a non-reactive tracer
injected with the reagent can be use to evaluate the degree
of mixing between water sources with the reaction zone
and to assess the overall rate of reaction.
As noted previously, modeling studies may provide a
supplementary line of evidence to assess the sensitivity of
the immobilized contaminant to changes in ground-water
chemistry. This type of analysis provides an indirect means
to gauge contaminant response, as well as a means to as-
sess the impact of other ground-water characteristics that
may not be practically assessed via direct measurements.
However, as previously noted, the degree of uncertainty
in model predictions will be constrained by the accuracy
of reaction expressions and parameter inputs to represent
aquifer conditions within the plume. Ultimately, these tests
may require more explicit analysis of the microbial com-
munity that supports existing conditions or mediates future
changes to redox conditions within the boundary of the
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plume. Approaches to identify and quantify active microbial
communities within the subsurface are discussed below.
IMC.4 Microbial Community Characterization
Microbiological evidence to support the natural attenua-
tion of inorganic contaminants as a remedial alternative
in ground water involves the characterization of microbial
population size, diversity, composition, physiological and
genetic/phylogenic traits. This section addresses tech-
niques used for the characterization of subsurface microbial
communities. In addition, some microbiological sampling
practices for the assessment of ground water are discussed
and emerging methodologies are identified.
IIIC.4.1 Standard and Emerging Techniques
Analyses focused on the composition and diversity of
bacterial community structures cannot rely on traditional
microbiological procedures alone. This is especially critical
in subsurface ecological systems because the vast majority
of the microbial communities that reside in that environment
have not been cultivated using culture-dependent methods
(Amann et al., 1995). Although the importance of emerging
molecular approaches in subsurface microbiology will be
stressed later in this section, a successful program will likely
include a combination of culture-independent methods as
well as traditional cultivation strategies. Often, molecular
tools are applied to pure culture isolates harvested from
defined media. In other words, traditional selective and
enrichment techniques are used to develop specific micro-
bial communities to be further characterized with molecular
monitoring (i.e., genes coding for 16S rRNAs [16S rDNA]
to identify potential gene expressions). Since standard
microbiological techniques are readily available in most
environmental laboratories, only a brief discussion of them
is provided here (see also Table 3.7).
• Most probable-number (MPN) technique is a standard
methodology used to estimate the number of specific
physiological types of bacteria. Usually, a modified
basal medium is amended with a carbon substrate and
electron acceptors for total heterotrophic aerobes or
anaerobes, denitrifying, iron-reducing, sulfate-reduc-
ing, and methanogenic bacteria (Fedorak et al., 1987;
Lovley, 1991; Mahne and Tiedje, 1995; Chapelle et
al., 2002; Tanner, 1989). A three-tube dilution series
is often used to provide a 10-fold dilution. The same
regiment is applied to semi-solid media to provide plate
counts.
• Acridine orange direct count (AODC) is the most ver-
satile technique for yielding a count of the total intact
cells without differentiation for viability (Ghiorse and
Balkwill, 1983). Differential staining for live vs. dead
can be used via the Bac-light™ method (Loyd and
Anthony, 1995).
• Phospholipid ester-linked fatty acids (PLFA) is a popu-
lar assay used to identifty "biomarkers" to provide a
quantitative insight into three important attributes of
microbial communities including viable biomass, com-
munity structure, and metabolic activity (Lehman et al.,
1995). An estimation of non-viable populations can be
accompanied through the measurement of diglyceride
fatty acids (DGFA). Both assays are independent of
the bias inherent in classical culturing techniques pro-
viding a more accurate estimation of in-situ microbial
populations. The lipid biomarker analysis is, however,
incapable of identifying every microbial species in an
environmental sample because many species contain
over-lapping PFLA.
• Community-level physiological profile (CLPP) can be
carried out using Biolog-GN plates (Biolog, Inc., Hay-
ward, CA). This "phenotypic fingerprinting" assay is
useful for screening bacterial isolates and consortia to
establish correlation between their activity and composi-
tion.
• rRNAa can be used for the determination of microbial
biomass (Loyd and Anthony, 1995).
Table 3.7 Standard and emerging techniques for micro-
bial community characterization.
Method
Most probable-number
(MPN)
Acridine orange direct
counts (AODC)
Phospholipid ester-linked
fatty acids (PLFA)
Community-level
physiological profile (CLPP)
Type of Information
Enumeration, Cultural
Enumeration,
Morphological Cultural
Enumeration,
Biochemical
Physiological
Microcosms are routinely prepared using subsurface
cores and ground-water samples for the characteriza-
tion of microbial communities. Recently, there has been
an increased interest in the use microcosms to perform
molecular community fingerprinting such as denaturing
gradient gel electrophoresis due to the generation of a
sufficient cell mass.
IIIC.4.2 Molecular Characterization
Within the last decade, a variety of culture-independent
genetic analyses have been used to complement traditional
culture-dependent methods (enrichment and isolation).
Many of these molecular biological methods rely on 16S
rDNA sequences, including in-situ hybridization (Amann
et al., 1990; Amann et al., 1995), direct amplification of
16S rDNA, and additional analysis using community "DNA
fingerprinting" such as temperature gradient gel electropho-
resis (TGGE) (Felske et al., 1998), denaturing gradient gel
electrophoresis (DGGE) (Muyzer et al. 1993), restriction
fragment length polymorphism (RFLP) (Martinez-Murcis
et al., 1995), single-strand conformation polymorphism
(SSCP) (Lee et al., 1996), terminal RFLP (Clement et al.,
1998), or 16S rDNA cloning-sequencing (Wise et al., 1997).
To date, most of the results obtained by using molecular
70
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techniques have been provided by cloning-sequencing of
16S rDNA fragments. Although 16S rDNA cloning-se-
quencing is successful on the reconnaissance of microbial
diversity by detecting infrequent sequences from various
habitats, thereby avoiding limitations of traditional cultiva-
tion techniques, it is time-consuming and problematic for
multiple sample analysis. Since the cloning approach
cannot provide an immediate overview of the community
structure, many environmental laboratories apply DGGE
to detect population shifts. DGGE can be used for simul-
taneous analysis of multiple samples obtained at various
time intervals to detect microbial community changes; an
advantageous feature in studying microbial ecology and
MNA.
These methods characterize differing aspects of the sub-
surface microbial community, which may be used alone or
in combination to further delineate the impact of microbial
processes on ground-water chemistry. DGGE provides a
simple approach to obtaining profiles of microbial communi-
ties and identifying temporal and spatial variations which
occur in response to various environmental conditions
(Muyzeret al., 1993). It is also possible to infer the phylog-
eny of community members by DMA sequence analysis of
re-amplified fragments, after they are excised from the gel,
where bands corresponding to each microorganism can be
separated through DGGE. Fluorescent in-situ hybridiza-
tion (FISH) provides a powerful tool for directly studying
organisms within the environment by providing information
on cell morphology, phylogenic affiliation and the ability to
quantify organisms (Amann et al.,1990; Amann et al., 1995).
During the last few years numerous efforts have been made
to increase the sensitivity of FISH, including multi-labeled
polynucleotide probes (Pernthaler et al., 2002). The RFLP
approach involves electrophoretic analysis where DMA is
detected with probes after Southern blotting. RFLP is ap-
plied broadly since it has a good predictive power and can
rapidly identify phylogenic relatedness of clusters of very
closely related or even identical strains. It can screen large
numbers of isolates to identify a much smaller subset of
representative types to be resolved by 16S rRNA sequenc-
ing. Although RFLP generates a complex set of rDNA
bands that can be used to group closely related strains, it
does not provide the distance between strains that are not
closely related, as does rRNA sequencing. In contrast, the
T-RFLP approach uses restriction enzymes, coupled with
PCR, in which only fragments containing a fluorescent
tag are detected. The use of T-RFLP is advantageous as
a rapid screening tool and does not require culturing or a
genetic database.
IIIC.4.3 Sampling Considerations
Hydrogeologic conditions are of overriding significance
in designing and conducting subsurface microbiological
sampling programs because the selection of equipment and
the location of sampling points are necessarily subject to
site-specific conditions. With regard to the type of sample
(core or water), core samples provide more information in
defining the horizontal and vertical distribution of microbes
even though they are intrinsically disruptive and prohibit
repeated sampling at the same location. Ground-water
samples, on the other hand, can be obtained from the same
well repeatedly but may not quantitatively or qualitatively
reflect conditions in the aquifer. Since there are uncertain-
ties that mandate caution in the extrapolation of information
obtained from either type of sample, a thorough character-
ization may require both water and core samples.
In studies addressing the origin and nature of subsurface
microbes, an important consideration is the extent organ-
isms cultured or manipulated in the laboratory represent
the intrinsic microbial community. Therefore, microbiologists
are challenged when collecting not only representative but
also microbially uncontaminated samples. To minimize
contamination, after a core is obtained using strict aseptic
methods, care should be taken not to disturb or contaminate
the sample. Processing should be performed as quickly
as possible under anaerobic and aseptic conditions while
in the field. Surface layers of the core should be scraped
away using a sterile sampling devise and discarded so that
only the center of the core is packaged in doubled sterile
sample bags. The portion used for DNA/PLFA analyses
must be rapidly frozen with liquid nitrogen and stored at
-70 L°C. Another portion should be flushed with inert
gases (N2 or Ar), sealed in canning jars and placed inside
cans containing oxygen-scavenging catalyst packets (Gas
Pak; BBL Inc., Franklin Lakes, N.J.), and stored at 4 °C for
microbial counts and cultural techniques for analysis within
24 hours. Frozen (-70 °C) genomic DNA can be extracted
from core sample using an UltraClean soil DNA kit (MoBio.
Solana, CA).
Ground-water samples will be unfiltered for microbial counts
and cultural techniques and filtered for DNA/PLFA analysis.
Community analysis involving "DNA fingerprinting" requires
ultrafiltration of ground water (50 to 100 L) using inorganic
(Anodisc™, Whatman) filters (0.2 urn pore size). The filter
is placed in sterile bag and rapidly freeze dried with liquid
nitrogen. Since PCR bias can be introduced, particularly
during cell lysis and PCR amplification, a physical method
such as bead mill homogenization should be used to ef-
fectively lyse all cell types, including those that are most
recalcitrant to physical and enzymatic treatments (More
etal., 1994).
MIC.5 Implications for Natural Attenuation
Assessment
Ultimately, the extent and degree of site characterization
to define the redox status of the aquifer will represent a
balance between technical information needs and the
cost associated with the different proposed data collec-
tion or evaluation schemes. The primary objective of site
characterization is to identify the mechanism leading to
contaminant attenuation at a given site. Emphasis of this
characterization effort should be given to direct measure-
ments of ground water conditions and solid phase char-
acteristics of the aquifer that result from these conditions.
Measurements and/or tests conducted with subsurface
samples retrieved within the zones where attenuation oc-
curs will provide the most direct means to evaluate on-going
reaction processes. This knowledge will guide approaches
71
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to assess the capacity of the aquifer to sustain contaminant
attenuation within the plume and to evaluate the long-
term stability of immobilized contaminants. Evaluations
conducted on subsurface samples also have the benefit
of implicitly incorporating characteristics/factors of ground
water and aquifer solids that may be difficult to adequately
parameterize within analytical models.
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