&EPA
United States
Environmental Protection
Agency
Monitored Natural Attenuation
of Inorganic Contaminants in
Ground Water
Volume 3
Assessment for Radionuclides Including
Tritium, Radon, Strontium, Technetium,
Uranium, Iodine, Radium, Thorium,
Cesium, and Plutonium-Americium
Attenuation Processes - Reaction Times (irxn)
220Rn 222Rn 131|
55.6s 3.82 of 8d
Radioactive Decay
241Am »6Ra "Tc
12.3 y 432 y 1600 y 210000 y
- -++- solid phase transformation
»H^rn Prec'P'tatlon . solid-state diffusion .
adsorption co-precipitation ^
<2 S2 1 100 10,000
~ 2 days months years I 1 1 1 1 1—>
E _c 10 1'000 100,000
surface water
groundwater
Transport Processes - Hydraulic Residence Times (t trans)
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EPA/600/R-10/093
September 2010
Monitored Natural Attenuation of Inorganic
Contaminants in Ground Water
Volume 3
Assessment for Radionuclides Including
Tritium, Radon, Strontium, Technetium,
Uranium, Iodine, Radium, Thorium, Cesium,
and Plutonium-Americium
Edited by
Robert G. Ford
Land Remediation and Pollution Control Division
Cincinnati, Ohio 45268
and
Richard T. Wilkin
Ground Water and Ecosystems Restoration Division
Ada, Oklahoma 74820
Project Officer
Robert G. Ford
Land Remediation and Pollution Control Division
Cincinnati, Ohio 45268
National Risk Management Research Laboratory
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, Ohio 45268
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Notice
The U.S. Environmental Protection Agency through its Office of Research
and Development managed portions of the technical work described here under
EPA Contract No. 68-C-02-092 to Dynamac Corporation, Ada, Oklahoma (David
Burden, Project Officer) 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
environmental data and funded by the U.S. Environmental Protection Agency are
required to participate in the Agency Quality Assurance Program. This project did
not involve the collection or use of environmental data and, as such, did not require
a Quality Assurance Plan.
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Foreword
The U.S. Environmental Protection Agency is charged by Congress with protecting the Nation's land, air, and water
resources. Under a mandate of national environmental laws, the Agency strives to formulate and implement actions
leading to a compatible balance between human activities and the ability of natural systems to support and nurture
life. To meet this mandate, EPA's research program is providing data and technical support for solving environmen-
tal problems today and building a science knowledge base necessary to manage our ecological resources wisely,
understand how pollutants affect our health, and prevent or reduce environmental risks in the future.
The National Risk Management Research Laboratory (NRMRL) is the Agency's center for investigation of technological
and management approaches for preventing and reducing risks from pollution that threatens human health and
the environment. The focus of the Laboratory's research program is on methods and their cost-effectiveness for
prevention and control of pollution to air, land, water, and subsurface resources; protection of water quality in public
water systems; remediation of contaminated sites, sediments and ground water; prevention and control of indoor
air pollution; and restoration of ecosystems. NRMRL collaborates with both public and private sector partners to
foster technologies that reduce the cost of compliance and to anticipate emerging problems. NRMRLs research
provides solutions to environmental problems by: developing and promoting technologies that protect and improve
the environment; advancing scientific and engineering information to support regulatory and policy decisions; and
providing the technical support and information transfer to ensure implementation of environmental regulations and
strategies at the national, state, and community levels.
This publication has been produced as part of the Laboratory's strategic long-term research plan. It is published and
made available by EPA's Office of Research and Development to assist the user community and to link researchers
with their clients. Understanding site characterization to support the use of monitored natural attenuation (MNA) for
remediating inorganic contaminants in ground water is a major priority of research and technology transfer for the
U.S. Environmental Protection Agency's Office of Research and Development and the National Risk Management
Research Laboratory. This document provides technical recommendations regarding the development of conceptual
site models and site characterization approaches useful for evaluating the effectiveness of the natural attenuation
component of ground-water remedial actions. This document addresses natural attenuation processes and data
requirements specific to selected radionuclides.
David G. Jewett, Acting Director
Ground Water and Ecosystems Restoration Division
National Risk Management Research Laboratory
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Contents
Notice ii
Foreword iii
Contents v
Figures ix
Tables x
Acknowledgements xi
Executive Summary xiii
Conceptual Background for Radionuclides 1
1.1 Background and Purpose 1
1.1.1 Document Organization 1
1.1.2 Purpose of Document 1
1.1.3 Policy Framework for Use of MNA 2
1.1.4 Applicable Regulatory Criteria 3
1.1.4.1 Radionuclide Standards 3
1.1.4.2 Use of Radionuclide Mass in Remediation 5
1.2 Contaminant Risk Reduction Processes 6
1.3 Tiered Analysis Approach to Site Characterization 9
1.3.1 Tier I 10
1.3.2 Tier II 10
1.3.3 Tier III 12
1.3.4 Tier IV 12
1.4 Incorporating Decay Phenomena into Descriptions of Subsurface Transport 13
1.4.1 Variable Solid-Liquid Partitioning for Parent and Daughter Radionuclides 13
1.4.1.1 Daughter In-growth 15
1.4.1.2 Data Sources and Determination of Solid-Liquid Partitioning 16
1.4.1.3 Influence of Alpha-Recoil on Daughter Solid-Liquid Partitioning 16
1.4.2 Colloid Generation and Transport 17
1.5 Site Characterization 17
1.5.1 Overview of Methods for Radionuclide Measurement 18
1.5.1.1 Radiometric Techniques 18
1.5.1.2 Mass-based Techniques 19
1.5.1.3 Radionuclide In-growth Corrections 19
1.5.2 Chemical and Redox Speciation 19
1.5.3 Multiple Sources for Radionuclide of Concern 19
1.5.3.1 Isotopic Composition for Radionuclide Source Discrimination 20
1.5.3.2 Identification of Progenitors 20
1.5.4 Procedures for Collection of Colloidal Radionuclide Forms 21
1.5.5 Aquifer Solids 22
Tritium 27
Occurrence and Distribution 27
Geochemistry and Attenuation Processes 27
Radioactive Decay 27
Adsorption 27
Site Characterization 27
Overview 27
Aqueous Measurements 27
Long-term Stability and Capacity 28
Tiered Analysis 28
References 29
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Radon 31
Occurrence and Distribution 31
Geochemistry and Attenuation Processes 31
Radioactive Decay 31
Aqueous Chemistry 31
Site Characterization 31
Overview 31
Aqueous Measurements 33
Solid Phase Measurements 33
Long-term Stability and Capacity 33
Tiered Analysis 34
References 35
Strontium 37
Occurrence and Distribution 37
Geochemistry and Attenuation Processes 37
Radioactive Decay 37
Aqueous Speciation 37
Solubility 38
Adsorption 38
Site Characterization 39
Overview 39
Aqueous Measurements 39
Solid Phase Measurements 40
Long-term Stability and Capacity 40
Tiered Analysis 40
References 41
Technetium 45
Occurrence and Distribution 45
Geochemistry and Attenuation Processes 45
Radioactive Decay 45
Aqueous Speciation 45
Solubility 45
Adsorption 46
Site Characterization 47
Overview 47
Aqueous Measurements 47
Solid Phase Measurements 48
Long-term Stability and Capacity 48
Tiered Analysis 49
References 50
Uranium 53
Occurrence and Distribution 53
Geochemistry and Attenuation Processes 53
Radioactive Decay 53
Aqueous Speciation 56
Solubility 57
Adsorption 59
Site Characterization 59
Overview 59
Aqueous Measurements 59
Solid Phase Measurements 61
Long-Term Stability and Capacity 61
Tiered Analysis 62
References 63
Iodine 69
Occurrence and Distribution 69
Geochemistry and Attenuation Processes 69
VI
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Radioactive Decay 69
Aqueous Speciation 69
Site Characterization 70
Overview 70
Aqueous Measurements 71
Solid Phase Measurements 72
Long-term Stability and Capacity 72
Tiered Analysis 73
References 74
Radium 79
Occurrence and Distribution 79
Geochemistry and Attenuation Processes 79
Aqueous Speciation 81
Solubility 81
Adsorption 82
Site Characterization 82
Aqueous Measurements 83
Solid Phase Measurements 83
Long-Term Stability and Capacity 84
Tiered Analysis 85
References 86
Thorium 91
Occurrence and Distribution 91
Geochemistry and Attenuation Processes 91
Radioactive Decay 91
Aqueous Speciation 93
Solubility 94
Adsorption 94
Site Characterization 95
Aqueous Measurements 96
Solid Phase Measurements 96
Long-term Stability and Capacity 97
Tiered Analysis 97
References 99
Cesium 103
Occurrence and Distribution 103
Geochemistry and Attenuation Processes 103
Radioactive Decay 103
Aqueous Speciation 103
Adsorption 103
Site Characterization 104
Overview 104
Aqueous Measurements 104
Solid Phase Measurements 104
Long-term Stability and Capacity 105
Tiered Analysis 105
References 107
Plutonium and Americium 111
Occurrence and Distribution 111
Geochemistry and Attenuation Processes 111
Radioactive Decay 111
Aqueous Speciation 113
Solubility 114
Adsorption 114
Site Characterization 114
Overview 114
Aqueous Measurements 114
Solid Phase Measurements 117
VII
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Long-term Stability and Capacity 117
Tiered Analysis 118
References 119
Appendix A - Radioactive Decay Processes 123
Modes of Radioactive Decay 123
Modes of Nuclear De-excitation (Following Decay) 124
Decay Chains 124
Units and Specific Activity 125
References 127
VIM
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Figures
Figure 1.1 Relative timescales of attenuation processes and fluid transport processes
referenced to characteristic reaction times and hydraulic residence times 6
Figure 1.2 Illustration of characteristic ground-water plumes that may develop for
radionuclides whose transport is not impeded by chemical reaction 7
Figure 1.3 Illustration of the importance of decay half-life and total activity on the viability
of radioactive decay as an attenuation process for contaminant remediation 8
Figure 1.4 Depiction of plume development for a radionuclide that possesses decay rate
slower than the rate of fluid transport, but is subject to sorption processes that
transfer it from ground water onto immobile aquifer solids 9
Figure 1.5 One-dimensional, decay-transport model describing tritium transport in an
homogeneous aquifer as a function of source characteristics 11
Figure 1.6 Illustration of four decay in-growth scenarios that may be encountered for
ground-water plumes contaminated with radionuclides 15
Figure 3.1 Potential radioactive decay paths that lead to the production of 222Rn 32
Figure 3.2 Daughter radioisotopes resulting from the radioactive decay of 222Rn 32
Figure 4.1 Decay series for 90Sr based on data from ICRR 37
Figure 4.2 Solubility and speciation of strontium as a function of pH and PCO2 with
comparison to the stability of calcite and siderite 38
Figure 5.1 Phase stability diagrams for technetium at 25°C as a function of pH 46
Figure 5.2 Phase stability diagrams for technetium at 25°C as a function of pH in the
presence of dissolved carbonate or sulfide 47
Figure 6.1 Decay series for 238U and 235U based on data from ICRP 55
Figure 6.2 Distribution of aqueous U(VI) species among oxyhydroxyl-, sulfate-, carbonate-,
and calcium carbonate-complexes as a function of ground-water chemistry. 56
Figure 6.3 Solubility of U(VI) as a function of pH at various levels of PCO2 and in the
presence of dissolved silica, calcium, and sodium 57
Figure 6.4 Solubility of U(VI) as a function of pH and in the presence of phosphate, sodium,
and calcium 58
Figure 6.5 Eh-pH stability diagrams for U at 25°C 58
Figure 7.1 Eh-pH diagram of dominant iodine aqueous species at 25°C 70
Figure 8.1 Production of 228Ra, 224Ra, and 223Ra 80
Figure 8.2 Decay series for 226Ra based on data from ICRP. 80
Figure 8.3 pH - log aSO42" diagrams showing stability fields for minerals in the
Ba-Ca-Ra-SO4-CO2-H2O system 81
Figure 9.1 Decay series for 229Th including parent 233U 92
Figure 9.2 Decay series for 234Th and 230Th including parents 238U and 234U 92
Figure 9.3 Decay series for 232Th including parent 236U and progeny 228Th 93
Figure 9.4 Solubility of ThO2(am) in (a) pure water and (b) 1 mM SO42' 95
Figure 9.5 Solubility of ThO2(am) over variable log PCO2 values (-4 to -2) and 1 mM SO42' 95
IX
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Figure 10.1 Decay reactions involving 135Cs and 137Cs 103
Figure 11.1 Decay series for 239Pu 112
Figure 11.2 Decay series for240Pu 112
Figure 11.3 Decay series for 241Pu, including daughter 241Am 113
Figure 11.4 Eh-pH stability diagram for Pu at 25 °C and PCO2 = 10-25atm 115
Figure 11.5 Solubility diagram for Am(l 11) at PCO2 = 10-25atm and PO2 = 0.2 atm 115
Figure A.1 Decay chain for 238U showing intermediate nuclides formed during series
transformation to stable 206Pb 126
Tables
Table 1.1 Examples of mass concentration equivalents to activity-based applicable
regulatory standards for select radionuclides in ground water 4
Table 1.2 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 14
Table 1.3 Illustration of potential decay paths from different progenitor sources
leading to production of either 239Pu or 240Pu within a plume 21
Table 2.1 Natural attenuation and mobilization pathways for tritium 27
Table 3.1 Natural attenuation and mobilization pathways for radon 33
Table 4.1 Natural attenuation and mobilization pathways for strontium 39
Table 5.1 Natural attenuation and mobilization pathways for technetium 48
Table 6.1 Illustration of potential decay paths from different progenitor sources
leading to production of uranium radioisotopes 54
Table 6.2 Natural attenuation and mobilization pathways for uranium 60
Table 7.1 Natural attenuation and mobilization pathways for iodine 71
Table 8.1 Natural attenuation and mobilization pathways for radium 82
Table 9.1 Natural attenuation and mobilization pathways for thorium 96
Table 10.1 Natural attenuation and mobilization pathways for cesium 104
Table 11.1 Illustration of potential decay paths from different progenitor sources
leading to production of americium and plutonium radioisotopes 111
Table 11.2 Natural attenuation and mobilization pathways for plutonium and
americium 116
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Acknowledgements
This document represents a collective work of many individuals with expertise in the policy and technical aspects of
selecting and implementing cleanup remedies at sites with contaminated ground water. Preparation of the various
components of this document was undertaken by personnel from the USEPA Office of Research and Development
(ORD), Office of Superfund Remediation and Technology Innovation (OSRTI), and Office of Radiation and Indoor Air
(ORIA), as well as technical experts whose participation was supported under USEPA Contract No. 68-C-02-092 to
Dynamac Corporation, Ada, Oklahoma (David Burden, Project Officer) through funds provided by ORIA and OSRTI.
Contributing authors are listed below:
Contributing Author
Robert G. Ford
Richard T. Wilkin
Daniel I. Kaplan
James E. Amonette
Patrick V. Brady
Paul M. Bertsch
Kenneth Lovelace
Stuart Walker
Ronald Wilhelm
Robert W. Puls
Craig Bethke
Douglas B. Kent
Affiliation
USEPA/ORD, National Risk Management Research Laboratory, Cincinnati, OH
45268
USEPA/ORD, National Risk Management Research Laboratory, Ada, OK 74820
Savannah River National Laboratory, Aiken, SC 29808
Pacific Northwest National Laboratory, Fundamental Science Directorate, Richland,
WA 99352
Sandia National Laboratories, Geochemistry Department (MS-0750), Albuquerque,
New Mexico 87185
University of Kentucky, Lexington, KY 40506
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
University of Illinois, Department of Geology, Urbana, IL 61801
U.S. Geological Survey, McKelvay Building (MS-465), Menlo Park, CA 94025
This document benefited from review by George Redden (Idaho National Laboratory), Andrew Sowder (EPRI), and
Sue Clark (Washington State University).
XI
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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 engi-
neering analysis informs the understanding, monitoring, predicting, and documenting of the natural processes. In
order to properly employ this remedy, the Environmental Protection Agency needs a strong scientific basis sup-
ported by appropriate research and site-specific monitoring implemented in accordance with the Agency's Quality
System. The purpose of this series of documents, collectively titled "Monitored Natural Attenuation of Inorganic
Contaminants in Ground Water," is to provide a technical resource for remedial site managers to define and assess
the potential for use of site-specific natural processes to play a role in the design of an overall remedial approach
to achieve cleanup objectives.
The current document represents the third 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 3, titled "Assessment for Radionuclides Including
Tritium, Radon, Strontium, Technetium, Uranium, Iodine, Radium, Thorium, Cesium, and Plutonium-Americium,"
consists of individual chapters that describe 1) the natural processes that may result in the attenuation of the listed
contaminants and 2) data requirements to be met during site characterization. Emphasis is placed on character-
ization of immobilization and/or radioactive decay processes that may control contaminant attenuation, as well as
technical approaches to assess performance characteristics of the MNA remedy. A tiered analysis approach is
presented to assist in organizing site characterization tasks in a manner designed to reduce uncertainty in remedy
selection while distributing costs to address four primary issues:
1. Demonstration of dissolved plume stability via radioactive decay and/or active contaminant removal from
ground water;
2. Determination of the rate and mechanism of attenuation by immobilization;
3. Determination of the long-term capacity for attenuation and stability of immobilized contaminants; and
4. Design of performance monitoring program, including defining triggers for assessing MNA failure, and
establishing a contingency plan.
Where feasible, Agency-approved analytical protocols currently implemented for waste site characterization are
identified, along with modifications that may be warranted to help insure the quality of site-specific data. In situ-
ations where Agency methods or protocols are unavailable, recommendations are made based on review of the
existing technical literature. It is anticipated that future updates to these recommendations may be warranted with
increased experience in the successful application of MNA as part of a ground-water remedy and the development
of new analytical protocols.
This document is limited to evaluations performed in porous-media settings. Detailed discussion of performance
monitoring system design in fractured rock, karst, and other such highly heterogeneous settings is beyond the
scope of this document. Ground water and contaminants often move preferentially through discrete pathways (e.g.,
solution channels, fractures, and joints) in these settings. Existing techniques may be incapable of fully 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 contaminants 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 expectations, or
the selection and implementation of MNA. This document does not supercede any guidance. It is intended for
use as a technical reference in conjunction with other documents, including OSWER Directive 9200.4-17P, "Use
of Monitored Natural Attenuation at Superfund, RCRA Corrective Action, and Underground Storage Tank Sites"
(http://www.epa.gov/swerust1/directiv/d9200417.pdf).
xiii
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Conceptual Background for Radionuclides
1.1 Background and Purpose
1.1.1 Document Organization
The purpose of this document is to provide a framework
for assessing the potential application of monitored natural
attenuation (MNA) as part of the remedy for inorganic con-
taminant plumes in ground water. It is organized as part of
three volumes that provide: Volume 1 - a general overview
of the framework and technical requirements for application
of MNA to inorganic contaminant plumes (USEPA, 2007a);
Volume 2 - contaminant-specific discussions addressing
potential attenuation processes and site characterization
requirements for non-radionuclides (USEPA, 2007b), and
Volume 3 - contaminant-specific discussions addressing
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 detailed overviews
in this volume are described below.
The radionuclide contaminants selected for this document
include: americium, cesium, iodine, plutonium, radium, ra-
don, technetium, thorium, tritium, strontium, and uranium.
The selection of these contaminants was based on two cri-
teria. First, a selected element had to be one of high priority
to the site remediation or risk assessment activities of the
USEPA (USEPA, 1993a; USEPA, 2002a; USEPA, 2006a;
USEPA, 2007c). Second, selection was based on chemical
behavior considering chemical traits such as: toxicity, cat-
ions, anions, conservatively transported, non-conservatively
transported, and redox sensitive elements (USEPA, 1999b;
USEPA, 2004a). 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.
7.7.2 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 contaminants' 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).
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
1 The term "inorganic contaminants" is used in this document as a ge-
neric term for metals and metalloids (such as arsenic); and also refers
to radiologic as well as non-radiologic isotopes.
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 contaminants.
The conceptual model provides the basis for assessing potential reme-
dial 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.
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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-
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 cannot be predicted directly
from the hydraulic gradient, and existing techniques may
not be capable of identifying the pathway along which
contaminated ground water 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.
1.1.3 Policy Framework for Use of MNA
The term monitored natural attenuation (MNA) is used in
this document when referring to a method of remediation.
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-specific remediation objectives within a
time frame that is reasonable compared to that
offered by other more active methods. The 'natu-
ral attenuation processes' that are at work in
such a remediation approach include a variety of
physical, chemical, or biological processes that,
under favorable conditions, 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; dilution; sorption;
volatilization; radioactive decay; and chemical or
biological stabilization, transformation, or destruc-
tion of contaminants. (USEPA, 1999c, page 3.)
Even though several physical, chemical, and biological
processes are included in the above definition, the"! 999
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 unac-
ceptable to the overseeing regulatory authority.
Therefore, sites where the contaminant plumes
are no longer increasing in extent, or are shrink-
ing, would be the most appropriate candidates
for MNA remedies." (USEPA, 1999c, page 18.)
Control of contaminant sources is also an important aspect
of EPA's policy. The actual policy language is given below:
"Control of source materials is the most effec-
tive means of ensuring the timely attainment of
remediation objectives. EPA, therefore, expects
that source control measures will be evaluated
for all contaminated sites and that source con-
trol 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 reme-
dial approach"), while collecting additional data
to determine 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
inorganic contaminants should be known at a particular
site, and the stability of the process should be evaluated
and shown to be irreversible. 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 (including
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 radionuclides (e.g., Am-241 and Np-237
from Pu-241) may be 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 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, or bioavailability such
that they pose a relatively low level of risk. There-
fore, natural attenuation of inorganic contaminants
is most applicable to sites where immobilization
or radioactive decay is demonstrated to be in
effect and the process/mechanism is irreversible.
(USEPA, 1999c, pages 8-9.)
1. 1.4 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 §300.430(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 stan-
dards include federal maximum contaminant levels
(MCLs) and/or non-zero maximum contaminant
level goals (MCLGs) established under the Safe
Drinking Water Act (SDWA), or mo re stringent state
drinking water standards. Other regulations may
also be ARARs as provided in CERCLA §121(d)
1.1.4.1 Radionuclide Standards
Current MCLs for radionuclides are set at 4 mrem/yr for the
sum of the doses from beta particles and photon emitters,
15 pCi/L for gross alpha particle activity (including 226Ra,
but excluding uranium and radon), and 5 pCi/L combined
for 226Ra and 228Ra. The current MCLs for beta emitters
specify that MCLs are to be calculated based upon an
3 When a contaminant is associated with a solid phase, it is usually
not known if the contaminant is precipitated as a three-dimensional
molecular 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.
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.
-------
annual dose equivalent of 4 mrem to the total body or any
internal organ. It is further specified that the calculation is
to be performed on the basis of a 2 liter per day drinking
water intake using the 168 hours data listed in "Maximum
Permissible Body Burdens and Maximum Permissible
Concentrations of Radionuclides in Air or Water for
Occupational Exposure," NBS Handbook 69 as amended
August 1963, U.S. Department of Commerce (U.S. DOC,
1963). These calculations have been done for most beta
emitters and published as part of the EPA guidance "Use
of Uranium Drinking Water Standards under 40 CFR 141
and 40 CFR 192 as Remediation Goals for Groundwater
at CERCLA sites" (OSWER No. 9283.1-14, November 6,
2001). This guidance also includes a list of radionuclides
addressed by the gross alpha MCL. The MCL for uranium
is 30 micrograms per liter (ug/L). Two isotopes of uranium
are also addressed by ground-water standards under the
Uranium Mill Tailings Radiation Control Act (UMTRCA). The
concentration limit for the combined level of 234U and 238U is
30pCi/L. Relevant standards for some of the radionuclides
discussed in this document are shown in Table 1.1.
Table 1.1 Examples of mass concentration equivalents to activity-based standards for select radionuclides in
ground water. Drinking water MCLs apply to total element concentration rather than specific radioactive
isotopes except where indicated. Fact Sheets with summary information on the radionuclides in this list
are available at http://www.epa.sov/superfund/health/contaminants/radiation/nuclides.htm.
Radionuclide
Contaminant
Americium
Cesium
Tritium
Iodine
Plutonium
Radium
Strontium
Technetium
Thorium
Uranium
241 Am
137Cs
3H
129|
238Pu
239Pu
240pu
241 pu
242Pu
244pu
226Ra
228Ra
90Sr
"Tc
228Th
229Th
230Th
232Th
234(J
238(J
Current MCLa or UMTRCA
(pCi/L)
15
200
20,000
1
15
15
15
(27 RBL)b
15
15
5C
5C
8
900
15
15
15
15
30d
30d
Mass Equiv to MCL, UMTRCA, or
RBL (ug/L)
0.0000044
0.0000023
0.0000021
0.0057
0.00000088
0.00024
0.000066
0.00000026
0.0038
0.85
0.0000051
0.000000018
0.000000059
0.053
0.000000018
0.000071
0.00074
140
30d
30d
a Federal Register, Vol. 65, No. 236, December 2, 2000; MCL is 4 mrem/yr to the whole body or an organ, combined from all beta and photon emit-
ters; MCL is 15 pd/L, with the concentration level combined for all alpha emitters, except radon and uranium.
b Risk Based Limits calculated for 30-year exposure duration and 1 x 1&6 risk. These were calculated using equation 11 in Risk Assessment Guid-
ance for Superfund (RAGS): Volume I: Human Health Evaluation Manual (Part B, Development of Risk-based Preliminary Remediation Goals),
(page 37). The equations were adjusted to account for radioactive decay.
c MCL is 5 pd/L combined for Ra-226 and Ra-228
d Federal Register, Vol. 42, No. 65, March 2, 2000, Rules and Regulations; MCL standard is 30 \ig/L for uranium; UMTRCA ground-water standard
is 30 pd/Lcombined for U-234 and U-238.
-------
1.1.4.2 Use of Radionuclide Mass in Remediation
Typically units of decay rate instead of mass are used to
quantify the concentration of radioactive material in con-
taminated environmental media because the carcinogenic
risks of exposure to radioactively-contaminated materials
are related more to the decay rate of the material than to
its mass.5 Generally, this convention is used due to the
short half-lives of many of the radionuclides commonly
encountered at contaminated sites. As examples for the
decay-equivalent for a given radionuclide mass, one gram of
226Ra has a decay rate (activity) of 3.7x1010 transformations
(also referred to as disintegrations) per second, while one
gram of 137Cs has a decay rate of 3.2x1012 transformations
per second. Except for long-lived nuclides with low specific
activities such as 238U, the energy emitted by the radioac-
tive material during radioactive decay and absorbed by
exposed biological tissue is the key driver of health effects
from exposure to most radionuclides. In addition, radioac-
tive materials may be detected and quantified by the type
of radiation emitted and number of disintegrations (per unit
time). For these reasons, the concentration of radioactive
material in water is typically expressed in units of activity
or decay rate, pCi/L.
Most of the radionuclide MCLs and UMTRCA ground-water
standards are presented in the traditional units of pCi/L.
Mass units, however, provide insight and information into
treatment selection, treatment compatibility, and treatment
efficiency, particularly for remedial actions involving mixed
waste. For example, remediation goals expressed in mass
are important for designing and evaluating treatment tech-
nologies such as soil separation, pump and treat, as well as
subsurface barriers. In addition, transport models in which
solid-liquid partitioning of the radionuclide is described
are developed using mass-action reaction expressions.
Radionuclide mass concentration is typically the required
data input for these models. Typically units for expressing
mass in environmental media for soil and water are mg/kg
and mg/L, respectively. These mass units also can be
expressed as parts per million (ppm) for soil and water,
which is equivalent to mg/kg and mg/L. MCLs in pCi/L
may be converted to their mass equivalent in mg/L, by the
following equations:
MCL(mg/L) = 2.8x 10'15 * A * T1/2 * MCL(pCi/L)
where 2.8x 10~15 for water is a conversion factor, A is the
radionuclide atomic weight in g/mole, and T1/2 is the radio-
nuclide half-life in years. Most radionuclides of concern for
site cleanups have half-lives ranging from a few years to
10,000 years. At MCL levels, the corresponding masses
of most radionuclides represent extremely small values.
One important issue associated with using mass to charac-
terize the quantities of radioactive material in the environ-
ment is that many elements, such as uranium, have several
isotopes of the same element (See examples in Table 1.1.).
For example, if one were to perform atomic absorption
5 Discussions of radioactive decay phenomena and applicable units of
measurement are provided in Appendix A.
analysis of a water sample, and it revealed the presence of
1 mg/kg of uranium, there would be no way of knowing how
much of the uranium in the sample is 238U, 234U, or 235U, all
of which are present in the environment naturally and due to
anthropogenic activities. While the potential human health
and ecological effects of uranium from its chemical toxicity
are impacted by the total mass of the element, its potential
for human health and ecological effects from its radioactivity
will depend on the specific isotopes of uranium present,
which could vary depending on whether one is dealing with
naturally-occurring uranium or uranium that may have been
enriched in 235U as part of the uranium fuel cycle or part
of weapons production. It is also important to note that
the same mass of each uranium isotope has significantly
different levels of radioactivity. A mass of 1 mg/kg of 238U
has an activity of 0.33 pCi/g, while the same mass of 235U
has 2.1 pCi/g and 234U has 6,200 pCi/g.
Also, many radioactive elements are present in the envi-
ronment along with their stable counterpart. One example
is potassium, which occurs naturally in the environment,
ranging from 0.1 to 1% in limestone to 3.5% in granite.
In addition, a typical 70 kg adult contains 130 g of potas-
sium. A very small fraction (0.01%) of this potassium is
the naturally-occurring radioactive isotope 40K. If one were
to measure the amount of 40K in soil and assume that 40K
made up all of the elemental potassium then the total mass
of this element would be underestimated by 10,000 fold.
Since the potential adverse effects of radioactive material
are due to the energy released following radioactive decay,
measurement of elemental mass present, e.g., total K by
atomic emission spectroscopy, may not accurately repre-
sent the amount of radioactivity present and, therefore, its
potential radiotoxicity. However, use of mass spectrometry
for discrimination of the various isotopes of a given element
may avoid this situation, since it would then be possible to
convert isotopic mass concentration to activity using the
decay half-life of a radioisotope.
Conversely, the measurement of the radioisotope activity
will be a misrepresentation of the total mass of the given
element, particularly for the majority of elements that have
non-radioactive isotopes which may be present in much
larger quantities on a mass basis. Accordingly, activity
should not be used alone to determine or tailor the treatment
required for remediation technologies, since technologies
typically rely on chemical and/or physical processes that
are sensitive to or driven by mass or concentration. For
example, to design and implement a treatment technology
for radioactive strontium (i.e., 90Sr), it would be necessary
to know the total mass of all stable (i.e., 84Sr, 86Sr, 87Sr,
88Sr) and radioactive isotopes of strontium in ground
water. The same considerations would be necessary for
other ground-water treatment technologies for dissolved
concentrations of elements and their isotopic forms. For
example in a pump and treat ground-water extraction
system that utilizes ion exchange (chemical separation)
or reverse osmosis (physical separation), chemical mass
measurements would be used to determine the amount
and type of reactant materials, exchange capacity and
-------
effectiveness (USEPA, 1996). Much the same can be
said for immobilization or reduction technologies such as
chemical solidification/stabilization treatability studies or
treatments (USEPA, 2000b). Also, mass measurements are
important in the determination of partition coefficients, Kd,
values that are often employed in risk assessment modeling
and remediation calculations. Kd values are expressed in
mass units for the inorganic elements and isotopes (USEPA,
1999b; USEPA, 2004a). The values of Kd are assumed to
be the same for all isotopic forms of the element.
In summary, given that risk or exposure is the basis for
remedial actions, mass measurements are often required
for determining, designing and selecting a remediation
technology. This contrasts with the need for radiation spe-
cific isotopic measurements required in risk and exposure
analysis. Users should note the different applications and
perspectives with their corresponding measurement units
of mass and activity.
1.2 Contaminant Risk Reduction Processes
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.' For radionuclides, contaminant attenuation
that results in mass loss or decreased mobility may occur
either via radioactive decay or immobilization. In general,
the development of a stable or shrinking contaminant
plume will depend on the relative rate for the immobilization
reaction(s) versus the rate of ground-water flow through
the aquifer. This concept is illustrated in Figure 1.1 in
which documented ranges for characteristic timescales of
radioactive decay and several potential immobilization reac-
tions are shown relative to commonly observed residence
times for a parcel of water in ground water (and surface
water) systems. Discussion of potential immobilization
processes and the types of site characterization data to
support identification of this attenuation process within a
ground-water contaminant plume is provided in Volume 1,
Section II of this series of documents (USEPA, 2007a). In
general, immobilization may occur as a result of precipita-
tion, co-precipitation, and/or adsorption reactions in which
the contaminant chemically reacts with dissolved and/
or solid components within the aquifer. For many of the
radionuclide contaminants discussed in this volume, there
are viable processes that may result in attenuation via im-
mobilization within the subsurface. More specific discussion
of immobilization mechanisms is provided in the individual
contaminant chapters. However, several of the radionuclide
contaminants discussed in this volume will remain mobile
in ground water due to their inherent chemical properties.
In these situations, radioactive decay may provide the only
viable mechanism for mass loss of the contaminant.
220Rn
55.6s
^
exchange-
adsorption
4
+ ^V
8 £ §
8 E «
4
Attenuation Processes - Reaction Times (!„„)
222Rn «ii Radioactive Decay 3H 241Am 22«Ra 99Tc
3.82 d 8d 12.3 y 432 y 1600 y 210000 y
^
. .... solid phase transformation
precipitation solid-state diffusion .
1 100 10,000
10 1,000 100,000
^
surface water *"
groundwater
Transport Processes - Hydraulic Residence Times (i trans)
Figure 1.1 Relative timescales of attenuation processes and fluid transport processes referenced to characteristic
reaction times and hydraulic residence times. For radioactive decay, representative radionuclides with a
range of decay half-lives (s = seconds, d = days, y = years) are shown. Representative time scales for
several processes that result in contaminant immobilization are also shown.
-------
For radionuclides with relatively short decay half-lives, the
development of a stable or shrinking ground-water plume
may occur if the rate of decay is greater than the rate of
ground-water transport. As illustrated in Figure 1.2 for
radionuclides whose transport within the aquifer is not im-
peded due to chemical reaction, the ground-water plume
may shrink, remain invariant (i.e., stable), or expand in size.
Plume shrinkage may occur if radioactive decay is fast rela-
tive to the velocity of ground-water flow (e.g., 222Rn (radon),
3.82 day half-life). Where the rates of decay and fluid trans-
port are relatively similar, development of a stable plume
may occur (e.g., 3H (tritium), 12.3 year half-life). Finally, for
long-lived radionuclides that may remain mobile (e.g., "Tc
(technetium), 2.13x105 year half-life), plume expansion is
anticipated to occur. Estimated differences in the relative
time scales to achieve cleanup levels for radionuclides
addressed in this volume are illustrated in Figure 1.3. The
trends shown in Figure 1.3 are based solely on the result
of first-order radioactive decay (see Appendix) of a finite
radionuclide activity ignoring the influence of ground-water
transport on the distribution of contaminant mass/activity
in space and time. There are several technical issues
that also factor into this evaluation, including the relative
magnitude (i.e., total activity of radionuclide) and rate of
release of the radionuclide from the source term(s) that
contribute to plume development (Figure 1.3c), as well as
the chemical, radiological and/or toxicological characteris-
tics of the decay product(s). For example, the complexity
introduced by the in-growth of a radioactive daughter with
different chemical properties would require simultaneous
tracking of an overlapping plume that could behave very
differently from that of the parent. However, the purpose of
these illustrations is to draw attention to the large dispar-
ity in radioactive decay rate for the radionuclides that are
encountered at contaminated sites. Further discussion of
these issues is provided later in this document.
Contaminant
Source
Original Plume
Boundary
Original
Regulated Plume
Boundary
Non-Regulated
Dissolved
Plume
Figure 1.2 Illustration of characteristic ground-water plumes that may develop for radionuclides whose transport is
not impeded by chemical reaction: 1) stable plume - similar rates of fluid transport and decay, 2) shrink-
ing plume - decay rate faster than fluid transport, and 3) expanding plume - decay rate slower than fluid
transport. Regulated mobile plume refers to that portion of the plume where contaminant mass/activ-
ity exceeds Risk-based orARAR criterion. The symbol "i" refers to the characteristic time for transport
(* trans) and radioactive decay (T ).
-------
II
JO Q
§,§>
ID 9
o: 1
s!
-------
Immobilized
'Solid-phase1
Plume
'Immobile' plume represents contaminant
mass sorbed onto aquifer solids at any point
in time
Future scenarios for evolution of 'immobile'
plume:
• Declines in mass & spatial
distribution due to radioactive decay
• Remains invariant in mass & spatial
distribution
• Evolves to new state that serves
as source for development of new
dissolved plume caused by:
1) Radioactive decay produces
more mobile daughter produces)
2) Changes in ground-water
chemistry cause re-mobilization
Figure 1.4 Depiction of plume development for a radionuclide that possesses decay rate slower than the rate of
fluid transport (i(rans > idecf>y), but is subject to sorption processes that transfer it from ground water onto
immobile aquifer solids (i.e., retardation factor, R> 1). The dissolved plume shrinks leaving behind an
immobile plume bound to aquifer solids.
Evaluating the overall success of natural attenuation for
remediation of radionuclides will require demonstrating that
the rate and capacity for contaminant attenuation meets
regulatory objectives (including time frame) and, in addi-
tion, that contaminant immobilization is sustainable to the
extent that future health risks are reduced to acceptable
limits. The latter requirement necessitates identifying the
chemical speciation of the contaminant partitioned to the
solid phase. This information is critical towards design-
ing laboratory tests and reaction transport models used
to assess and project potential for re-mobilization of the
radionuclide. An overview explaining the types of immo-
bilization mechanisms and the respective susceptibility to
re-mobilization for each scenario is provided in Volume 1
of this series of documents (USEPA, 2007a) along with
general procedures for assessing the susceptibility for con-
taminant re-mobilization. Specific discussions of radioactive
decay phenomena, relevant immobilization processes and
procedures for assessing the potential for re-mobilization
are discussed within the individual contaminant chapters
later in this volume.
1.3 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 ap-
proach seeks to progressively reduce uncertainty as site-
specific data are collected. The decision-making approach
-------
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
contaminants in ground water. The fourth tier is included
to emphasize the importance of determining appropriate
parameters for long-term performance monitoring, once
MNA has been selected as part of the remedy. Data col-
lection 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 immobilization 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
attenuate 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.
7.3.7 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 radionuclides in which radioactive decay is
anticipated to provide the primary mode of attenuation, it
is recommended that decay calculations (with or without in-
corporation of physical transport) be conducted to evaluate
whether regulatory objectives can be met in an appropriate
time frame given knowledge of source term characteristics
(e.g., total activity and release rate) and/or radionuclide
activities within the plume relative to points of compliance.
For contaminants in which sorption onto aquifer solids is
the most feasible attenuation process, an additional ob-
jective 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.
An example illustrating the influence of total radionuclide
activity on plume dynamics is shown Figure 1.5. Two sce-
narios are shown for a tritium plume that results from two
different periods of release from a source area. In the left
panel, the release of tritium into ground water at a fixed
solution activity of 50,000 pCi/L occurred for a period of
12 months. For the hydrologic conditions specified in this
scenario (Figure 1.5), insufficient attenuation would be
projected based on radioactive decay alone. The activ-
ity of tritium in the plume centerline exceeds the MCL as
the plume maximum passes the point of compliance. In
contrast, if source control had been implemented such that
release only occurred for a period of 1 month, it might be
anticipated that radioactive decay may be sufficient to meet
a cleanup objective such as an MCL for ground water. In this
case, the maximum tritium activity is at or below the MCL
as the plume maximum passes the point of compliance.
This type of screening analysis is important to conduct
early in the site characterization effort for radionuclides for
which attenuation by immobilization is not significant. This
illustration also points to the importance of understanding
source characteristics both in terms of total contaminant
mass/activity as well as rates of release.
Determination of contaminant sorption onto aquifer solids
could be supported through the collection of aquifer cores
coincident with the locations of ground-water data collec-
tion 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 was provided in Volume 1 (USEPA,
2007a; Figure 1.2 in Section IC.1). Ultimately, sites that
demonstrate ground-water plume expansion and a lack
of contaminant sorption would be eliminated from further
consideration of MNA as part of the cleanup remedy.
7.3.2 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 timeframe that is reasonable compared
to other remedial alternatives, (see USEPA, 1999c, pages
19-21, for a discussion of "reasonable timeframe 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
specific mechanism(s) controlling contaminant attenuation.
An estimate of attenuation rates for long-lived radionuclides
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 recom-
mended that these estimates be based as much as possible
on field measurements rather than modeling 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 was provided in Volume 1,
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,
10
-------
60000
50000 -
(A) Source Pulse = 12 months
(B) Source Pulse = 1 month
Plume Centerline
^^~ 6 months
—•—12 months
—•—18 months
A 24 months
• 30 months
• 36 months
Horizontal Distance from Source (m)
Horizontal Distance from Source (m)
Figure 1.5 One-dimensional, decay-transport model (Clement, 2001; Quezada et al., 2004; and Srinivasan
and Clement, 2008 a & b) describing tritium transport in an homogeneous aquifer as a function
of source characteristics: (A) 12-month release from source at 50,000 pd/L and (B) 1-month
release from source at 50,000 pd/L. Assumptions: ground-water seepage velocity = 100 m/y,
longitudinal dispersivity = 0.11 m,3H decay half-life = 12.35 y insignificant influence from source term
decay.
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
Volume 1, 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 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., U(VI) vs. U(IV)] as
well as specific associations with chemical constituents in
aquifer solids (e.g., precipitation of uranium oxide/silicate
vs. adsorption of U(VI) to iron oxides). Evaluations of the
subsurface microbiology may be necessary in situations
where biotic processes play a direct or indirect role in gov-
erning 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.
11
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7.3.3 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 immobilization
is the primary attenuation process. For immobilized con-
taminants, factors to consider relative to the long-term
stability of the attenuated contaminant include 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 adsorption is typi-
cally 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
water. 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., ambi-
ent 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 utility
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 conducted through the previous
Tiers indicates that the aquifer within the plume boundar-
ies 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.
7.3.4 Tier IV
The objective under Tier IV analysis is to develop a moni-
toring program to assess long-term performance of the
MNA remedy and identify alternative remedies that could
be implemented for situations where changes in site condi-
tions 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
12
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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
oxides orsulfide 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 man-
ner that allows the ability to eliminate sites at intermediate
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 1.2. 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 hydrol-
ogy, 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 objectives 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 selec-
tion. 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 characteriza-
tion and general factors to consider for implementation of
a long-term performance monitoring program. These top-
ics are addressed at this juncture to allow greater focus to
discussions later in this volume pertaining specifically to
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.
1.4 Incorporating Decay Phenomena into
Descriptions of Subsurface Transport
The use of models to describe ground-water flow and con-
taminant transport for non-radionuclides was provided in
Volume 1 of this document (Section I.D. in USEPA, 2007a;
see also USEPA, 1996). In general, the types of models
and supporting data needed as input data to describe
contaminant transport of long-lived radionuclides will
be similar to that needed for non-radionuclides in which
chemical reactions control mass transfer within the aquifer.
However, for short-lived radionuclides or for radionuclides
in which the energy released during decay can impact
reaction conditions, additional model constructs and/or
input data may be needed. Models have been developed
that can simultaneously describe contaminant transport for
a parent radionuclide and daughter products that display
variable degrees of sorption affinity for aquifer solids (e.g.,
van Genuchten, 1985; Srinivasan and Clement, 2008 a & b).
This approach has application to short-lived radionuclides,
since the influence of radioactive decay on the distribution
of radionuclide activity/mass can be projected within the
plume boundary (e.g., Figure 1.5). In these model applica-
tions, sorption to aquifer solids is typically represented by
a linear sorption coefficient or "Kd" (see discussion below).
This approach is most likely applicable to radionuclides and/
or daughter products in which adsorption or ion exchange
is the dominant mechanism for solid-liquid partitioning.
However, it should be noted that these model applications
may be limited both by the inadequacy of a linear sorption
coefficient to describe contaminant partitioning (e.g., Bethke
and Brady, 2000; Zhu, 2003) and the failure to account for
contaminant precipitation reactions that may accompany
changes in redox chemistry and/or major ion fluxes within
the plume. Specific discussions on approaches to monitor
and model contaminant solid-liquid partitioning are provided
in the individual contaminant chapters later in this volume.
1.4.1 Variable Solid-Liquid Partitioning for Parent
and Daughter Radionuclides
As described previously, radioactive decay processes result
in the formation of a new radionuclide that may have signifi-
cantly different chemical properties. Thus, while radioactive
decay will result in mass-loss of the parent radionuclide, a
more mobile daughter radionuclide may result. As an ex-
ample, 226Ra (typically associated with solids) decay to 222Rn
(a dissolved gas) presents a situation where the parent and
daughter isotopes display very different reaction-transport
characteristics in ground water. This may be an undesir-
able situation if the daughter radionuclide presents a similar
13
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Table 1.2 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
contaminant removal
from ground water
Ground-water flow direction (calculation of hydraulic gradients); aquifer
hydrostratigraphy
Contaminant concentrations in ground water and aquifer solids
Evaluation of potential for plume expansion based on estimation of activity/
mass removal via radioactive decay compared to ground-water transport
velocity in aquifer
General ground-water chemistry
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
attenuation
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
heterogeneity
Select monitoring parameters to assess consistency in hydrology,
attenuation 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
14
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or greater radiological or chemical risk. For example, the
amount of radiation being released can actually rise over
time as successive radioactive decay products undergo
decay. Ultimately, the degree to which one needs to be
concerned about this issue will depend on the number and
total activity of radionuclide contaminants that are present
in a particular plume. Given this information, one can make
use of radioactive decay relationships to project increases
in decay products and activity, or in-growth, with or without
consideration of ground-water transport.
1.4.1.1 Daughter In-growth
Depending on the parent radionuclide of concern, there
are four general scenarios that may be encountered in
which unstable and/or stable nuclides may be produced:
1) unstable parent decays to a stable daughter (e.g., 137Cs
decay to stable 137Ba), 2) unstable parent and unstable
daughter have similar half-life (e.g., 227Th decay to 223Ra),
3) unstable parent has much longer half-life than unstable
daughter (e.g., 226Ra decay to 222Rn), and 4) unstable par-
ent has much shorter half-life than unstable daughter (e.g.,
241 Am decay to 237Np). The changes in parent and daughter
activities/concentrations along with total activity in ground
water are illustrated for these four scenarios in Figure 1.6.
It can be seen that the production of daughter products can
influence plume composition, potential radiological risks,
and the dimensions of the plume if the daughter product
displays a radiological or chemical risk and transport char-
acteristics different from that of the parent radionuclide.
Parent (unstable) - Daughter (stable)
Parent (unstable) - Daughter (unstable)
" ~ T1/2,d
combined activity
^original radionuclide
decay product
period of
ingrowth
transient equilibrium
Parent (unstable) - Daughter (unstable)
T"l/2,p >> "
combined activity
Parent (unstable) - Daughter (unstable)
Tia.p«'
'•5
IB
original radionuclide
f
ecular
| equilibrium
original radionuclide
,decay product
combined activity
time
period of ingrowth
I
PI
period of
ingrowth
no equilibrium
Figure 1.6 Illustration of four decay in-growth scenarios that may be encountered for ground-water plumes contami-
nated with radionuclides. Illustrations were derived from the USEPA website - http://www. epa. gov/radia-
tion/understand/equilibrium.html. T1/2 = decay half-life of parent radionulide, T1/2 d = decay half-life of
daughter radionuclide.
15
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1.4.1.2 Data Sources and Determination of Solid-
Liquid Partitioning
As described in Section IIIB.3 in Volume 1 (USEPA, 2007a),
there are several different approaches to describing solid-
liquid partitioning in transport models. The simplest ap-
proach to modeling solid-liquid partitioning during ground-
water flow is to use a linear sorption coefficient, or Kd, that
represents contaminant uptake as a linear function of the
total concentration in solution. The use and limitations of
this approach are provided in various publications (e.g.,
USEPA, 1999a; Bethke and Brady, 2000; Davis and Curtis,
2003; Zhu, 2003). One significant limitation to use of the Kd
approach is that the sorption coefficient is generally devel-
oped for (and only applicable to) a fixed set of ground-water
chemical conditions. The chemical conditions in ground
water affect the speciation of radionuclides in solution (e.g.,
complexation with dissolved carbonate), surface chemical
properties of potential sorbents (e.g., charging behavior
of oxyhydroxide minerals), and the types and tendency
for chemical bond formation onto aquifer solids (e.g., pH
dependence of cation sorption). Thus, even if a Kd was
developed using well-preserved aquifer solids and sampled
ground water from the site being characterized, chemical
conditions may evolve within the plume to the extent that
the developed Kd relationship may no longer provide a
valid representation of radionuclide partitioning. From this
perspective, use of a Kd obtained from literature reports or
derived from site-specific tests should be done so following
critical evaluation of the compatibility between test condi-
tions and subsurface characteristics with the portion(s) of
the contaminant plume for which contaminant attenuation is
being assessed. Under many situations, selection of a Kd
can single-handedly determine the results of contaminant
attenuation calculations.
There are available compilations of sorption coefficients
for the radionuclides addressed in this volume (USEPA,
1999 a & b; USEPA, 2004a). In these technical reviews,
contaminant-specific partitioning coefficients were derived
from published data sets for a wide range of soils. While
these data may have limited applicability for a specific site,
they do provide useful context for evaluating the relative
mobility of various radionuclides that might be present in
a ground-water plume. In contrast, there are more limited
compilations describing radionuclide partitioning to various
minerals in aquifer solids (e.g., iron oxyhydroxides, clay
minerals such as montmorillonite) that are based on more
detailed mechanistic descriptions of solid-liquid partition-
ing (e.g., surface complexation models; see Section MB in
Volume 1). However, the applicability of these models is
usually limited to describe radionuclide transport in contami-
nant plumes due to the uncertainty in input parameters to
describe mineral surface charging and/or the potential role
of multiple mineral components participating in solid-liquid
partitioning along transport pathways. Hybrid modeling
approaches have been developed in which variable char-
acteristics of sorbent phase(s) along transport pathways
is addressed via collection of site-specific data, while al-
lowing the ability to incorporate the influence of aqueous
radionuclide speciation on the partitioning reaction (e.g.,
Davis and Curtis, 2003).
In situations where radionuclide precipitation reactions
may control attenuation, verification of the accuracy and
completeness of solubility reactions included in the ther-
mochemical database is critical for model applications.
There have been significant efforts over the past several
years to review and update thermochemical databases
available to describe radionuclide precipitation in ground-
water systems. Examples of these efforts include those by
the Nuclear Energy Agency (Guillaumont et al., 2003) and
the Paul Scherrer Institute Laboratory for Waste Manage-
ment (Thoenen et al., 2004) to critically evaluate existing
published thermochemical data and provide the technical
rationale for selection of specific constants based on the
reliability of published methods and results in the technical
literature. It is recommended that any modeling effort that
incorporates descriptions of radionuclide precipitation as
an attenuation mechanism make use of these technically
reviewed databases.
Ultimately, the effort expended into the development of
reactive-transport models will likely be governed by the
type of attenuation processes anticipated to control radio-
nuclide transport. It is recommended that the selection
of appropriate modeling approaches be supported with
the collection of site-specific data that verify the primary
attenuation mechanisms for each radionuclide in a given
ground-water plume (i.e., as set forth in recommendations
for Tier II evaluation efforts).
1.4.1.3 Influence of Alpha-Recoil on Daughter
Solid-Liquid Partitioning
In general, solid-liquid partitioning for radionuclides
addressed in this volume will be governed by properties
of the radionuclide and sorbent material, as well as the
overall chemical conditions in ground water. As previously
discussed, radioactive decay may result in daughter
products that possess sorption characteristics significantly
different than the parent nuclide. Radioactive decay of
sorbed radionuclides may also alter properties of the solid
to which they are partitioned. For some radionuclides, the
energy released during radioactive decay is sufficient to
either eject the daughter element from the solid structure
(Kigoshi, 1971) or cause damage to the host solid
(Fleischer, 1980). For the latter situation, the damage to
the solid structure increases the susceptibility to dissolution
(Eyal and Fleischer, 1985).
The potential influence of alpha-recoil is exemplified by
the behavior of 238U and 234U solid-solution partitioning in
ground-water systems in which 234U becomes enriched
in ground water (Ivanovich, 1994). For this decay chain,
238U decay produces 234Th along with emission of an alpha
particle of sufficient energy to cause ejection of 234Th into
ground water. Subsequent serial decay of 234Th (24.1 day
half-life) to 234Pa (6.7 hour half-life) and ultimately 234U re-
sults in an elevated activity/concentration of 234U relative to
what would be anticipated based strictly on the solid-liquid
partitioning for uranium and/or thorium. Additional examples
16
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of decay chains that may produce recoil effects include
those for 228Th, 229Th, and 226Ra (Sun and Semkow, 1998).
Radionuclide transport models have been developed that
incorporate the influence of alpha-recoil on the solid-liquid
partitioning of parent and daughter radionuclides (Tricca et
al., 2001; Maheretal., 2006). However, this phenomenon is
not routinely recognized as a contributing factor in ground-
water plumes at contaminated sites.
In general, the impact of this process is difficult to character-
ize in an aquifer and may play a minor role in contaminant
plumes with concentrations that greatly exceed natural
levels. However, given that mass-equivalent regulatory
standards for many of the radionuclides addressed in this
volume are quite low, the potential importance of alpha-
recoil events should be qualitatively evaluated before ruling
out the significance of this process. In other instances,
consideration of alpha-recoil processes may be more rel-
evant from the standpoint of explaining anomalous analyti-
cal results and/or selecting/tailoring an analytical method.
7.4.2 Colloid Generation and Transport
An overview of colloid generation and transport in ground
water was provided in Section IIA.2 in Volume 1 (USEPA,
2007a). In that discussion, emphasis was placed on col-
loidal solids acting as carriers of inorganic contaminants
within a ground-water plume. The types of colloids dis-
cussed included minerals and/or organic compounds native
to the aquifer solids or formed from components in waste
streams transported as part of the ground-water plume. It
has been proposed that these colloid types be referred to
as "pseudo-colloids" when describing radionuclide transport
(Kim, 1991). The characteristics of carrier colloids and tech-
nical issues important for site characterization presented
in Volume 1 also apply to radinonuclide transport in this
volume. For the radionuclides addressed in this volume,
an additional source of colloids may be derived from the
radionuclide under certain chemical conditions. These mo-
bile solids are referred to as radiocolloids or "real-colloids"
(Kim, 1991), which can be derived from the polymerization
and precipitation of the radionuclide to form solids with size
dimensions on the order of nanometers. An example of
this phenomenon is the formation of fine-grained thorium
oxide solids possessing surface charge characteristics that
maintain their mobility in porous media (e.g., Yun et al.,
2006). A conceptual framework for understanding colloid
stability and transport in ground water has been developed
(Degueldre et al., 2000), and it is recommended that this be
consulted as a point of reference for assessing the potential
importance of this process at a given site.
Conflicting reports of the importance of colloidal transport
for a given site exist in the literature (e.g., Savannah River
Site - Kaplan et al., 1994 and Dai et al., 2002; Los Alamos
National Laboratory - Penrose et al., 1990 and Marty et
al., 1997). Review of these findings point to the critical
need to 1) insure use and documentation of appropri-
ate well design, construction, and screen development
and/or re-development procedures, 2) employ sampling and
analysis protocols that avoid the generation of analytical
artifacts, and 3) develop a comprehensive knowledge of
the potential sources of the contaminant along transport
pathways throughout the plume (e.g., Dai et al., 2002; see
Section 1.5.3 below). These observations suggest that
colloidal transport is often invoked to explain apparent en-
hanced contaminant transport in place of direct observation
of colloids along plausible transport pathways. In general, it
appears that the common approach of identifying the pres-
ence of colloidal matter through comparison of filtered and
unfiltered water samples will not be a reliable approach for
confirming this transport mechanism for radionuclides. It is
recommended that characterization of the chemical com-
position and structural identity of the purported colloids be
determined in order to evaluate whether the identified solids
are likely derived from the aquifer solids along transport
pathways or if they are an artifact resulting from improper
well development and/or sampling protocols. There is evi-
dence that colloid mobilization may be short-lived within a
given aquifer due to the high surface area and reactivity of
colloidal materials (e.g., Miekeley et al., 1992). This is of
particular importance for the development of contaminant
transport models that may be used to project plume ex-
pansion or contraction. At present, the level of uncertainty
relative to the capability to directly observe and model
colloidal transport in ground water should be considered
to be high. A brief review of sampling and analysis tech-
nologies for the collection and analysis of colloidal phases
in ground water is provided below (Section 1.5.4) for sites
where colloidal transport appears to play a critical role in
contaminant transport.
1.5 Site Characterization
The objective of characterization efforts at a site where MNA
is being considered as a component of the ground-water
remedy is to evaluate the performance characteristics of
existing conditions within the aquifer to achieve cleanup
goals. Unlike engineered remedies where certain basic
performance characteristics are generally understood, se-
lection of MNA depends on developing detailed understand-
ing of the active attenuation mechanisms within the plume
and evaluating how changes in ground-water chemistry
may impact the rate, capacity and long-term stability of
contaminant attenuation. In order to develop a mechanistic
understanding of the attenuation process(es) active within
a plume, acquisition of characterization data describing
hydrologic conditions, ground-water chemistry, contaminant
speciation, and factors controlling solid-liquid partitioning is
needed. Review of the types of required data, approaches
to obtain these data, and the approaches to make use of
these data to assess the feasibility of MNA selection was
provided in Volume 1 of this document (USEPA, 2007a).
Specifically, discussion within Volume 1 covered charac-
terization of site hydrology (Section IMA), characterization
of ground-water chemistry including aqueous contaminant
speciation (Section IIIB.1), and characterization of aquifer
solids and the product(s) of contaminant immobilization
reactions (Section NIB.2). These characterization tasks
also apply to the radionuclides discussed within this volume
and specific recommendations are included in the individual
17
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contaminant chapters. The objective of this section is to
highlight characteristics and analytical methodologies that
are unique to contaminants subject to radioactive decay.
1.5.1 Overview of Methods for Radionuclide
Measurement
As with stable elements, detection of radionuclides can
be achieved using mass-based techniques (e.g., mass
spectrometery) or the measurement of the interaction of
various sources of electromagnetic radiation with the target
radionuclide or chemical complexes it may form with other
chemical constituents (e.g., absorption of radiation in the
ultraviolet or visible spectrum). In general, the sensitivity
of chemical methods based on absorption or emission of
electromagnetic radiation is not sufficient to achieve detec-
tion levels comparable to mass-equivalent regulatory levels
for many of the radionuclides of concern. Exceptions to this
generalization are provided in the individual contaminant
chapters included later in this volume, where applicable.
Absorption- or emission-based chemical methods are also
limited to measurement of total radionuclide abundance
and are unable to differentiate between isotopes of a given
element. However, the property of radioactive decay pro-
vides another approach to uniquely identify and quantify
radionuclides. Radioactive decay results in the formation of
a daughter element that may be unique to the decay event,
and it also may result in the production of energetic particles
(e.g., alpha particles) and/or radiation (e.g., gamma radia-
tion) possessing energies that are also specific to the decay
process (e.g., as reviewed in USEPA, 2006b). Thus, detec-
tion of these inherent products of the decay process can
be used as an approach to uniquely identify and quantify
the parent radionuclide. Since radioactive decay reactions
follow known mechanisms with fixed reaction rates, it is
generally feasible to measure the products of decay and
calculate the activity/mass of the parent radionuclide that
was necessary to produce the observed product activity/
mass. As discussed in USEPA (2006b), these methods
may be used for the purpose of detecting the presence of a
particular radionuclide (i.e., screening methodologies) and
the quantification of the activity- or mass-based concentra-
tion in the sampled medium.
There are some limitations to the use of radioactive decay
as the basis for radionuclide detection. A notable example
is the inability to distinguish 239Pu and 240Pu using alpha par-
ticle detection, since decay of both radionuclides produces
alpha particles with energies that are unresolved employing
available energy detection devices (e.g., Dai et al., 2002).
The following discussion provides a brief overview of the
more commonly used analytical methods with published
standard approaches, as well as more recently developed
mass-based methodologies that are seeing greater applica-
tion for characterization of ground water with radionuclide
contamination (e.g., Hou and Roos, 2008). It should be
noted that many radionuclides of concern have low specific
activities (e.g., 238U or 235U), and their measurement may
be more appropriately addressed using mass-based meth-
odologies (See Section 1.5.1.2 and individual radionuclide
chapters for additional information.)
1.5.1.1 Radiometric Techniques
A recent compilation of radiometric methods used for site
characterization to support environmental restoration efforts
has been published by the USEPA (2007c). Section 6.0
of this compilation of standard analytical methods (SAM;
Version 3.1; available at http://www.epa.gov/nhsrc/pubs/re-
portSAMOBOl07.pdf\ includes radiochemical methods that
have a high likelihood of assuring analytical consistency.
Criteria used to make this characterization included the
historical use of the referenced methods and the availability
of laboratory facilities across the nation with the capability
and capacity to conduct sample analyses. The methods
listed in this compilation include only radiometric methods
based on the detection and quantification of gamma rays,
alpha particles, or beta particles emitted during decay of
the targeted radionuclide or a common daughter product,
where the activity or mass of the radionuclide is derived
using known characteristics of the decay process. The
listed methods address analysis of both liquid and solid
samples and sample preparation procedures needed based
on requirements of the detection method. The methods can
be used for determination of qualitative (i.e., radionuclide
identification) and quantitative (i.e., radionuclide activity or
mass) information for the sample matrix.
A broader list of radiochemical methods is provided in the
Multi-Agency Radiological Laboratory Analytical Protocols
(MARLAP) Manual available at http://www.epa.gov/radia-
tion/marlap/manual.html. Analysis of liquid samples (or sol-
ids following dissolution) using the radiometric techniques
described above requires significant sample processing
1) to isolate radionuclides of interest from potentially inter-
fering matrix components, and 2) to present the sample
in a configuration that optimizes counting statistics (e.g.,
to minimize absorption of alpha or beta particles prior to
detection). A number of methods and techniques employed
to separate and purify radionuclides contained in environ-
mental samples are described in Chapter 14 (Separation
Techniques) of the MARLAP Manual posted at the website
listed above (USEPA, 2004b). Details on the analytical
requirements, potential interferences and performance
characteristics for the various radiometric techniques are
provided in Chapter 15 of the MARLAP Manual (Quantifica-
tion of Radionuclides; USEPA, 2004b). This document also
provides discussion of the use of liquid scintillation meth-
ods in which decay of a radionuclide excites a compound
("fluor") that produces fluorescent radiation.
It is recommended that site managers also consult a con-
cise compendium of methods for radionuclide detection
and quantification in soils and water, entitled "Inventory of
Radiological Methods" (USEPA, 2006b), which is available
at http://www.epa.gov/narel/IRM_Final.pdf. In this docu-
ment, a summary of nominal minimum detection limits for
the various radionuclide detection methods is provided
in Figure 2 (pg. 41) and a list of applicable methods for
the radionuclides discussed in this volume is provided in
Table 10 (pg. 42) of this publication (USEPA, 2006b).
18
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Recent developments in the design of analytical systems
that combine the steps of radionuclide isolation from the
sample matrix in-line with various radiometric detection
systems have facilitated the application of radiometric
measurements in the field. Grate et al. (2008) provide a
comprehensive review of the types of systems that have
been developed along with their performance character-
istics for detection and quantification of radionuclides in
ground water. In situ sensors for long-term monitoring ap-
plications without use of consumable reagents have also
been developed (e.g., Egorov et al., 2001) for the quanti-
fication of "tc and 90Sr in ground water. These analytical
platforms will help address the need to collect spatial and
temporal information on ground-water plume behavior.
For contaminants in which desired detection limits can be
achieved with short detector count times, these techniques
provide the means to adjust sampling activities in the field.
1.5.1.2 Mass-based Techniques
For very long-lived radionuclides (those with half-lives
over 10,000 years, e.g., 234/235/2331^ 239/24o/244pu_ 99jCi 129^
mass-based techniques may be faster and more sensi-
tive than nuclear-decay emission analyses. More recent
environmental applications include the use of mass-based
techniques for the analysis of nuclides with intermediate
half lives, including 90Sr, 226Ra, and ^'^Cs. In addition,
sample preparation for mass-based techniques can avoid
some of the radionuclide separation and purification steps
required for nuclear-decay emission analyses, providing
added savings in time and labor, which is particularly true for
alpha spectroscopy where sample self-adsorption concerns
can lead to extensive labor intensive sample preparation.
Becker (2003) and Lariviere et al. (2006) provide recent
reviews of mass spectrometry applications for the deter-
mination of radionuclide concentrations in environmental
samples. The sensitivity and mass-selectivity of these ap-
proaches, along with the ability to circumvent matrix and
isobaric interferences, have significantly increased the utility
of these methods. This can be further improved with the
use of on-line techniques for the separation and enrichment
of the targeted radionuclide from the sample matrix (e.g.,
Egorov et al., 2001; Miro and Hansen, 2006), similar to
approaches used for radiometric measurements. These
methods also provide the opportunity to simultaneously
measure the concentrations of a range of stable isotopes
that may provide information necessary to uniquely iden-
tify potential waste sources and/or ground-water transport
pathways within the aquifer.
1.5.1.3 Radionuclide In-growth Corrections
As previously discussed, the activity and mass of a radio-
nuclide continuously decreases in time due to radioactive
decay (in the absence of production from decay of pro-
genitors that may exist in contaminant source areas; see
Section 1.5.3.2). For short-lived radionuclides, this may
impact analytical results, since contaminant measurements
are typically conducted after a period of time has elapsed
since the time of sample collection. Since sample collection
is conducted to provide a snapshot of plume conditions at
a particular point in time, it is critical that analytical data
for radionuclides be adjusted to account for decay losses
during the interval between sample collection and analysis.
Specifically, there is no method of preservation that can halt
contaminant losses due to radioactive decay. These correc-
tions apply both for situations where the target radionuclide
is the parent isotope (i.e., decreasing activity with time) or
a daughter isotope produced from decay of other radionu-
clides that may be present in the sample (i.e., increasing
activity with time). An example of this latter phenomenon
was illustrated by Dai et al. (2002) where a fraction of 240Pu
in ground-water samples was produced from decay of
244Cm that was also present in the sample during storage.
Without correction for production of 240Pu during storage,
the activities of this radionuclide would have been biased
high, providing an inaccurate picture of the activity or con-
centration of 240Pu at the time of sampling from within the
plume. In this situation, 240Pu accumulates in the sample
with time due to its long half-life for radioactive decay. Due
to the relatively low regulatory benchmarks for activity- and
mass-based radionuclide concentrations in ground water,
radionuclide in-growth corrections may be important for
accurate descriptions of plume characteristics in time.
7.5.2 Chemical and Redox Speciation
Determination of the chemical speciation and redox state of
a radionuclide in ground water is critical for assessing the
factors controlling contaminant attenuation (or mobilization).
As a point of reference, uranium is typically more mobile in
its oxidized form [U(VI)j, although the uranyl cation may ad-
sorb onto mineral surfaces. Elevated alkalinity can suppress
adsorption of the uranyl cation through formation of soluble
carbonate complexes (Um et al., 2007). Confirmation of
the speciation of uranium for this situation would be best
achieved through confirmation of the predominant oxida-
tion state of uranium in solution. For uranium, it would be
possible to measure the relative proportion of U(VI) using
kinetic phosphorimetry, since only the oxidized species is
luminescent (Brina and Miller, 1992). Determination of the
fraction of uranyl complexed with carbonate would then
necessitate measurement of alkalinity (or inorganic carbon)
with calculation of carbonate species using a chemical spe-
ciation model. Approaches to determine either redox state
or solution speciation for specific radionuclides are provided
in individual contaminant chapters later in this volume,
where available. May et al. (2008) provide a recent review
of methodologies that have been used to determine aque-
ous speciation of long-lived radionuclides in environmental
samples, including evaluation of the analytical merits of the
various methods for determining contaminant speciation.
7.5.3 Multiple Sources for Radionuclide of
Concern
Characterization of radionuclide transport within a plume
is commonly evaluated through the installation and sam-
pling of multiple monitoring points within the aquifer. The
transport behavior of a specific radionuclide is then inferred
through measurement of the specific activity or mass con-
centration of the radionuclide in space and time relative
19
-------
to the measured velocity of ground-water flow through the
aquifer. In many cases, evaluation of the physicochemical
processes controlling transport of a specific radionuclide
may be based on assumptions about the characteristics of
fluid transport or the potential source(s) of the radionuclide.
Since knowledge of source term characteristics is critical to
assessing attenuation mechanisms and/or capacity within
the aquifer, proper identification of the apparent up gradi-
ent source of the radionuclide at any location within an
apparent plume is important to development of an accurate
conceptualization of contaminant transport. Two issues that
are important to consider relative to proper identification of
the contaminant source include: 1) differentiating whether
a single source term or multiple source terms contribute to
the contaminant mass/activity throughout the plume, and
2) determination of whether the specific radionuclide was
present in the initial source zone or whether it is a product
of decay of more mobile parent radionuclides moving within
the plume.
1.5.3.1 Isotopic Composition for Radionuclide
Source Discrimination
Ground-water plumes may be derived from one or multiple
sources at a site. In addition, large sites may have multiple
source areas that contribute to multiple plumes that may or
may not intersect at some location within the aquifer. Since
determination of the capacity of the aquifer to attenuate
the mass of contaminant within the plume is identified as
a critical component of site characterization (Tier III in Sec-
tion 1.3.3), information on the mass of contaminant being
transported along ground-water flow paths within the aquifer
needs to be determined. Radionuclide contaminants at a
site may be derived from anthropogenic (i.e., historical dis-
posal of process wastes) or natural sources. Comparison
of the distribution of isotopes for a given radionuclide near
suspected source areas and down gradient within a plume
provides an approach to determine whether a single or
multiple areas of contamination contribute to the mass of
contaminant within the aquifer where MNA is being evalu-
ated as a component of the ground-water remedy (e.g.,
Ketterer et al., 2004). This characterization approach can
be supplemented with analysis of 1) unregulated radio-
nuclides that may be characteristic of a particular waste
stream and transported with the contaminant of concern
(Brown et al., 2006; ruthenium as a tracer for different tech-
netium sources) or 2) the distribution of stable isotopes for
an element that is characteristic of the hydrologic source
of ground water observed at a given location within a
plume. As an example, studies examining the distribution
of radionuclides and stable isotopes have been conducted
at the Hanford Site in Richland, Washington to aid in dif-
ferentiating contaminant sources for ground-water plumes
at various locations throughout the site (Dresel et al., 2002;
Christensen et al., 2004; Christensen et al., 2007). This
information has been used to constrain interpretations
of ground-water transport and projections of the mass of
contaminants potentially being transported along critical
transport pathways.
For radionuclide contaminants that may be derived from
both anthropogenic and natural sources, determination of
the contaminant isotopic distribution may provide one line
of evidence for identifying the predominant contaminant
source. For anthropogenic sources, knowledge of the
manufacturing or utilization process may also be important
for situations where characteristics of the raw materials may
have varied over the lifetime of the process. For example,
the manufacture of targets for production of plutonium may
have made use of depleted or natural forms of uranium.
Uranium also provides an example of a radionuclide that
may have been present as a component of a waste stream
disposed on site or derived from aquifer solids due to leach-
ing reactions driven by movement of the plume through
the aquifer. For example, isotopic ratios of assy^sey or
236y:238y jn ground water or aquifer solids within the plume
may provide a signature for an anthropogenic source of this
radionuclide (e.g., Marsden et al., 2001; Howe et al., 2002).
This approach is applicable at sites where wastes may have
been derived from operation of nuclear reactors or the pro-
cessing of uranium fuels to support nuclear reactions. The
unstable isotope 236U does occur in nature (2.3415x107 yr
half-life), but only at ultra-trace concentrations with a
236y.238U atom ratjo of -| Q-14 (ZhaQ et a| -| 997) Jhe ratjo Qf
these isotopes is anticipated to be higher in wastes derived
from reactor operations due to the production of elevated
levels of 236U during neutron irradiation of 235U. In addition,
the ratio of assy^sey js anticipated to be higher in waste
materials derived from use of 235U due to enrichment of
this radioisotope relative to 238U in materials used in reactor
operation (relative natural abundance of 238U and 235U is
99.275% and 0.720%, respectively). Conversely, the ratio
of 235y:238y js anticipated to be lower in waste materials
with depleted levels of 235U due to extraction processes to
produce a material enriched in 235U for reaction operations
(Meinrath et al., 2003). As illustrated by these possible
scenarios, determination of isotopic ratios in environmental
samples from a contaminated site provides a potentially
important tool for determining contaminant sources and
tracking contaminant transport within an aquifer.
1.5.3.2 Identification of Progenitors
Radionuclides in a contaminant plume may be present as a
native component or as a daughter product from decay of
progenitor radionuclides within a waste stream. Knowledge
of process history for radionuclide production at a site and
source term composition can help in the identification of
possible progenitors. Since the progenitor and daughter
radionuclides may have differing transport properties, this
may complicate determination of the mechanism controlling
transport of the daughter radionuclide (i.e., contaminant of
concern) at different locations within the aquifer. Dai et al.
(2002) provide a useful example where the presence of
progenitors whose decay produces plutonium daughter
products could potentially lead to misidentification of the
controlling transport mechanism. In their field research,
these authors identified that 244Cm decay led to significant
production of 240Pu within a ground-water plume. The 244Cm
was present in a waste disposal area that was located along
20
-------
the path of an apparent single plume of plutonium within
the aquifer. In previous research, Kaplan et al. (1994) had
rationalized the apparently long transport distances for plu-
tonium as being due mobile colloids. Through examination
of the activities of plutonium isotopes and potential curium
and/or americium progenitors documented in records of
site activities and waste production, Dai et al. (2002) were
able to demonstrate that 240Pu observed at large distances
from the source area was actually derived from in-growth
from decay of 244Cm that was more mobile in ground water.
Examples of possible progenitors that could lead to in-
growth of 239Pu or 240Pu are illustrated in Table 1.3. The rela-
tive importance of progenitors for a given site will depend
on the types of radionuclides in various waste sources, the
total activity and half-life of the progenitors, and the rela-
tive mobility of the progenitors. As illustrated in Table 1.3,
progenitors with half-lives that are significantly shorter than
the daughter radionuclide of concern will be the ones that
are most likely to cause significant contributions from in-
growth. In addition, the mobility of the progenitor relative
to the chemical conditions within the plume may also need
to be evaluated as a part of the site characterization ef-
fort. The potential for contaminant in-growth during plume
transport points to the importance of understanding the
characteristics of potential waste sources relative to the
types of radionuclides that might be present in addition to
those that may be specifically targeted from a regulatory
perspective.
Table 1.3 Illustration of potential decay paths from different progenitor sources leading to production of either 239Pu
or 240Pu within a plume. Due to the relatively short half-life for the progenitor radioisotopes, appreciable
activities of 240Pu or 239Pu may result if appreciable activities of the progenitors are present in the plume.
Determination of possible decay paths to the target radionuclide was based on examination of the Chart
of Nuclides (http://www.nndc.bnlgov/chart/} maintained by the Brookhaven National Laboratory, National
Nuclear Data Center relative to possible decay paths based on decay modes identified in the Appendix
(EC = electron capture, p~ = electron emission, a = alpha decay). Decay half-life data were obtained us-
ing the WinChain program that provides electronic access to the ICRP38 Nuclear Decay Data Files (ICRP,
1983; Eckerman et al., 1994; m = minutes, h = hours, d = days, y = years). WinChain is a public domain
software application available for download from Oak Ridge National Laboratory
(http://ordose. ornl. gov/downloads. html).
Contaminant
Radionuclide
239Pu
240pu
Decay
Progenitor
239Np
243Cm
243Am (via 239Np)
240Np
244Cm
244Am (via 244Cm)
Decay
Mode
P-
a
a(p-)
P"
a
p-(a)
Progenitor
Decay Half-life
2.355 d
28.5 d
7380 y
65 m
18.11 y
10.1 h
7.5.4 Procedures for Collection of Colloidal
Radionuclide Forms
Determination of whether colloidal transport is a factor for
radionuclide transport in ground water will be predicated
on implementation of well installation, development, and
sampling protocols that avoid potential artifacts leading to
colloidal loss or production of colloidal material not pres-
ent within the plume. As previously discussed in Volume 1
(USEPA, 2007a; Sections IIA.2.1 and IIIB.1.1), improper
development of newly installed well screens and/or purging
at high volumetric rates during sampling can also lead to
the production of suspended solids that may mistakenly be
identified as mobile colloids within the aquifer. Well instal-
lation procedures may be a source of suspended solids
retrieved during well sampling. Two common types of solids
include fine-grained materials used in drilling fluids (e.g.,
bentonite) and colloidal-sized aquifer solids either dislodged
from the matrix or resulting from breakdown of larger-sized
matrix particles due to physical forces from the drilling ac-
tivity. Introduction of anthropogenic, fine-grained materials
may be avoided through the use of drilling methods that do
not require lubricants such as clay suspensions or through
the use of water as the drilling fluid. Aggressive develop-
ment of the well (e.g., surging and/or high pumping rates)
can be used to remove these types of solids from within
21
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the well screen and the portion of the aquifer in contact
with the screened interval. Another potential source of
artifactual solids retrieved during sampling may be due to
precipitated solids that accumulate on well-screen materi-
als and are physically dislodged by the sampling activity
(e.g., Houben, 2006). An example of this type of solid is the
accumulation of iron oxides that result from the oxidation
and precipitation of ferrous iron during natural or induced
intrusions of dissolved oxygen within the screened interval
of the aquifer. These types of solid accumulations may
be susceptible to being dislodged at the beginning of well
purging to establish stabilization for sampling. From this
perspective, it is recommended that the well screen be re-
developed, when feasible, prior to initiating low-flow purging
to establish stabilization of field parameters.
Another source for colloidal solids in the sampled water
may derive from solid precipitation reactions occurring at
the land surface during sample handling (e.g., Nilsson et
al., 2008). As illustrated by Dai et al. (2002), Buesseler et
al. (2003), and Hassellov et al. (2007), formation of colloidal
iron oxyhydroxides is a potential artifact when sampling
from ground-water plumes under iron-reducing conditions.
These authors have developed sampling procedures that
prevent exposure to air, and, therefore, reduce the potential
for rapid oxidation and precipitation of iron oxyhydroxides.
Characterization of the chemical composition or mineralogy
of recovered colloids provides one approach to assess the
potential for artifacts. For example, the presence of fine-
grained iron oxyhydroxides in ground water with elevated
ferrous iron concentrations [>10 mg/L Fe(ll)] is an indication
that iron oxidation-precipitation reactions may be occurring
during sample collection or processing. The use of proce-
dures such as cross-flow filtration may also be necessitated
to reliably isolate colloids from the ground-water sample
with minimal introduction of artifacts (e.g., Dai et al., 2002;
Hassellov et al., 2007). As shown in the literature (e.g.,
Baumann et al., 2006; Baik et al., 2007), colloidal fractions
that serve as carriers for radionuclide transport may reside
in multiple size fractions that would typically pass a 0.45 u,m
membrane filter. Thus, use of a conventional filtration pore
size such as 0.45 u,m will not provide a reasonable means
for differentiating between colloidal and truly dissolved
contaminant forms in sampled ground water.
In general, colloidal sampling procedures have not pro-
gressed to the point of being a routine practice during
ground-water sampling. Given that many of the regulatory
limits for radionuclides equate to extremely low mass-based
concentrations, there is a need to develop consensus ap-
proaches that reliably sample colloidal forms of contami-
nants from ground water. Ultimately, it is recommended
that ground-water sampling for the purpose of evaluating
the presence of mobile colloids be conducted using per-
manent monitoring points that have a clear record of well
construction (including description of granular solid materi-
als employed during drilling and screen placement within
the formation), procedures used for screen development,
and the data used to evaluate the adequacy of well devel-
opment and sampling procedures.
7.5.5 Aquifer Solids
The objectives and methodologies presented in
Sections NIB and NIC in Volume 1 (USEPA, 2007a) for
solid phase characterization to support evaluation of MNA
as a component of a ground-water remedy are directly
applicable to radionuclides. In general, characterization
of radionuclide speciation along with the determination
of abiotic or biotic solid components that participate in
contaminant immobilization represent the primary data
requirements. Many of the radionuclides discussed in this
volume can exist in multiple oxidation states in ground-water
systems (e.g., uranium, plutonium, americium, technetium).
Changes in oxidation state can dramatically alter transport
characteristics of these radionuclides. An example of
this phenomenon is the influence of reduction-oxidation
reactions that transform uranium between U(VI) and
U(IV) oxidation states. Different pathways for sorption or
precipitation exist for U(VI) and U(IV), thus identification of
the reaction mechanism controlling uranium immobilization
will, in part, depend on knowledge of the oxidation state
of uranium in aquifer solids. Knowledge of the controlling
immobilization mechanism will subsequently govern the
approach for assessing the capacity and stability of the
attenuation process.
An additional consideration for characterization of
radionuclide speciation in aquifer solids is the distribution
of isotopes of the element. Methods are available for
measuring isotope distributions in solid samples real-
time in the field (ITRC, 2006) or in the laboratory using
established procedures (USEPA, 2004b; USEPA, 2006b;
USEPA, 2007d). For some of the radionuclides addressed
in this document, there are no natural sources for aquifer
solids (e.g., 240Pu). However, several of the radionuclides
addressed in this document will have natural sources due
to the presence and decay of natural uranium (e.g., 234U,
230Th, 226Ra, and 222Rn from decay of 238U) or other naturally
occurring elements for which radioisotopic forms exist. In
addition, stable isotopic forms exist for some radionuclides
addressed in this document (e.g., strontium). The potential
presence of multiple isotopes of a given element within
the solid matrix points to the need to measure specific
isotopes during analyses that depend on mass-specific
detection. An example situation could be the application
of extraction-based approaches to determine either total
contaminant mass or chemical speciation within aquifer
solids (e.g., USEPA, 2007a, Section IIIB.2; Filgueiras et
al., 2002; Bacon and Davison, 2008). The distribution of
isotopes within the aquifer solids can also be used to verify
active immobilization reactions. It is anticipated that the
distribution of long-lived isotopes in the solid matrix will
reflect the isotopic distribution in ground water in portions
of the plume where immobilization is actively occurring
(Payne and Airey, 2006).
22
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26
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Tritium
Daniel I. Kaplan, Robert G. Ford
Occurrence and Distribution
Tritium (3H) is formed through natural and anthropogenic
processes. Tritium is formed in the upper atmosphere
through the interaction of fast neutrons with nitrogen
(Jacobs, 1968; MacKenzie, 2000). Natural background
levels have risen since weapons testing of nuclear devices
from background levels of 1 to 10 tritium units to several
hundred tritium units in the late 1960s (e.g., Egboka
et al., 1983). [Tritium unit (TU) denotes the number of
tritium atoms per 1018 atoms of hydrogen (Jacobs, 1968)
and is calculated by the expression, TU = 3H activity/
(1018* 1H activity).] Levels of 3H have since returned to near
background levels as a result of the atmospheric detonation
moratorium. Anthropogenic sources of 3H derive from the
production, use, and reprocessing of nuclear materials
(Jacobs, 1968) or from land disposal of commercial
products that incorporate 3H as a functional component
(Mutch and Mahony, 2008). Along with uranium, tritium is
the most common radioactive contaminant found in ground
water on U. S. Department of Energy (DOE) sites; identified
on 12 of the 18 DOE facilities (DOE, 1992).
Geochemistry and Attenuation Processes
Radioactive Decay
Hydrogen has three isotopes: stable protium (1H), stable
deuterium (2H), and radioactive tritium (3H). The relative
abundances of 1H, 2H, and 3H in natural water are 99.984,
0.016, and 1x10'15 percent, respectively (Freeze and
Cherry, 1979). Tritium has a half life for radioactive decay
of 12.3 years, and disintegrates into stable 3He by emission
of a beta particle. Tritium oxidizes rapidly to form tritiated
water, 1HO3H, and its distribution in nature is controlled by
the hydrologic cycle (Jacobs, 1968).
Adsorption
Tritium is generally considered not to sorb to aquifer solids
and is typically assigned a partition coefficient, Kd, of 0 ml_/g
to describe its partitioning to geological solids (McKinley and
Scholtis, 1993). All field studies indicate that tritium move-
ment is indistinguishable from water movement (USEPA,
1999). Some non-zero Kd values for tritium were reported
in the Kd compilation put together by Thibault et al. (1990).
These non-zero Kd values were as great as 0.1 mL/g, but
it is unclear what specific mechanism may have produced
these observed non-zero values. Contemporary evalua-
tions of 3H transport in ground water assume conservative
behavior for this constituent (e.g., Hu et al., 2008).
Site Characterization
Overview
Attenuation of 3H might be achieved through radioac-
tive decay (Table 2.1). Two factors that will dictate the
adequacy of attenuation via radioactive decay include the
rate of water transport and the total mass and release rate
of 3H into the subsurface plume. Evaluation of whether
radioactive decay is sufficient to achieve cleanup goals will
necessitate developing knowledge of the characteristics
of ground-water flow throughout the plume, as well as the
total activity/mass of 3H within the plume and entering from
uncontrolled source areas.
Table 2.1 Natural attenuation and mobilization path-
ways for tritium.
Attenuation
Processes
Radioactive
decay
Mobilization
Processes
Not applicable
Characterization
Approach
Determination of
ground-water velocity
along relevant
transport pathways
and contaminant mass
release rate from
source areas.
Aqueous Measurements
Relative to determination of the spatial distribution and
temporal variations in 3H concentration/activity within the
plume, methods for the preservation and analysis of 3H in
ground water are reviewed in USEPA (2006; Section 3.2.1
and Section 4, Table 10, respectively). As a low energy
beta-emitter, 3H is most commonly analyzed using liquid
scintillation radiometric methods. Possible interferences
to this method include the presence of other radioisotopes
emitting beta/alpha particles or gamma rays (e.g., 60Co, 88Y,
137Cs), as well as other constituents that cause coloration
of the water sample and potentially quench detection of
fluorescence from the scintillation reaction. Warwick et al.
(1999) have documented a simple distillation methodology
to minimize analytical bias from these potential interfer-
ences. For sites with buried contaminant sources in the
unsaturated zone, surface infiltration into the underlying
saturated aquifer or periodic saturation from water table
fluctuations may serve as a continuing source to a 3H
plume. Mapping out the locations and dimensions of these
unsaturated zone sources may prove critical to reliable
assessment of attenuation capacity within the aquifer. Field
studies have shown that determination of 3He/4He ratios
in soil gas from direct-push wells, via mass spectrometric
27
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detection in gas samples, may provide a useful technique to
characterize contaminant source areas (Olsen et al., 2001;
Peterson et al., 2002; Olsen et al., 2006). Since radioactive
decay of 3H produces 3He, elevated 3He/4He ratios provide
a potential signature of an elevated source of subsurface
3H. While this analytical approach is currently not applied
in a routine manner, it may provide critical information for
assessment of attenuation potential. In addition, measure-
ment of 3H, 3He, and 4He has been used to track sources
of water in the subsurface, including estimates of the age
of ground water (e.g., Egboka et al., 1983; Poreda et al.,
1988; Visser et al., 2007). For this purpose, it is important
to assess whether isolated sources of dissolved gases,
e.g., from volatile organic compounds, methane production,
and/or N2 production from denitrification, may influence
the partitioning of 3H, 3He, and 4He within ground-water
samples collected from well screens positioned along the
subsurface transport pathway being investigated (Visser
et al., 2009). This information can supplement character-
ization of the hydrologic system influencing contaminant
transport through the plume.
Long-term Stability and Capacity
Since immobilization is not an active mechanism for 3H
attenuation in ground water, assessment of long-term
stability is not a factor. The long-term capacity for 3H
attenuation within a plume will be dictated by the relative
rate of ground-water flow along relevant transport path-
ways compared to the rate for radioactive decay, given
a known flux of 3H entering the plume. Thus, a critical
factor for assessing the overall capacity of the aquifer for
attenuation will be evaluation of the mass flux of 3H moving
through the plume relative to the rate of water movement
through the aquifer. Radioactive decay may be sufficient
to prevent plume expansion, but this is not likely for sites
with an uncontrolled source of 3H entering the subsurface
and/or characteristic times for ground-water transport that
are significantly shorter than the half-life of radioactive
decay. Relative to 3H release from uncontrolled source
areas, knowledge of the total contaminant mass as well
as the rate and frequency of release into the saturated
aquifer needs to be developed. As an example, Taffet et
al. (1991) have shown that fluctuations in the water table
elevation due to periodic infiltration or recharge events can
result in periodic flushing of 3H from typically unsaturated
zones into the saturated aquifer. In order to make a reliable
assessment of the mass/activity flux of 3H into the plume,
it will be important to understand the characteristics of
the hydrogeologic system and the dynamics of water and
contaminant transfer from contaminant source areas into
the plume.
Tiered Analysis
Determination of the viability of 3H remediation in ground
water via monitored natural attenuation will depend upon
proper assessment of contaminant loading to the aquifer
relative to the velocity of ground-water travel along relevant
transport pathways to the point(s) of compliance. The
goal of site assessment is to demonstrate that radioactive
decay is adequate to meet cleanup goals given current and
projected hydrologic conditions for the site. The following
tiered analysis structure for site characterization provides
an approach to evaluate candidate sites and define the
potential limitations of MNA as part of a remedy for ground-
water cleanup.
Tier I. Site characterization under Tier I will involve dem-
onstration that the ground-water plume is static or shrink-
ing, has not reached compliance boundaries, and does
not impact existing water supplies. It is also important
at this stage of the site evaluation to determine source
term characteristics such as the inventory of contaminant
mass and the current and historical rate of release into the
subsurface. Acquisition of this information in combination
with identification of a stable plume provides justification
for proceeding to Tier II characterization efforts.
Tier II. Under Tier II, the apparent rate of attenuation is
determined. Estimates of a site attenuation rate(s) can be
assessed via a well transect along the ground-water flow
path. In addition, time-series data may be collected at
one or more monitoring points within the plume (USEPA,
2007; Section IIIA.5). Since radioactive decay will be the
dominant mechanism controlling 3H attenuation, determi-
nation of the velocity of ground-water flow along relevant
transport pathways will be critical to evaluating the potential
cleanup time frame. This information will allow assessment
of the relative timescales for contaminant attenuation and
fluid transport and determination of whether remediation
objectives can be met within the required regulatory time
frame. Determination of the adequacy of radioactive
decay to achieve cleanup goals will necessitate detailed
analysis of system hydrology relative to flow pathway(s),
flow velocity, and temporal variations in flow velocity and/or
direction within the boundaries of the plume. This informa-
tion, in combination with knowledge of contaminant source
location(s), tritium mass, and release characteristics, can
be employed to develop a decay-transport model to project
3H activity/concentration distribution throughout the plume
(e.g., see Figure 1.5). The demonstration of concurrence
between conceptual and mathematical models describing
tritium transport will entail development of site-specific
parameterization of ground-water flow along relevant
transport pathways.
Tier III. Under Tier III, it will be important to assess whether
the capacity of the system is adequate to sustain 3H attenu-
ation (e.g., prevent plume expansion) relative to the mass
of the contaminant being transported through the plume.
Two principal factors that may influence capacity include
changes in water transport and/or changes in 3H mass/
activity flux entering the plume. As identified by Taffet et al.
(1991), source release characteristics may be influenced
by the physical location of sources within the subsurface
relative to the ground-water table. Fluctuations in infiltra-
tion through shallow, unsaturated zones and/or water table
elevations within the aquifer due to variations in recharge
may lead to periodic increases in 3H release from con-
taminant source areas. Changes in land usage (including
at recharge zones, wetlands, and over the contaminated
28
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site) and/or ground-water withdrawals from the aquifer may
influence ground-water flow direction and velocity (USEPA,
2007; Section IMA), which in turn directly influences 3H travel
time. It is recommended that additional tritium transport
modeling be included to evaluate the impact of these vari-
ous scenarios to be assured that these perturbations do
not significantly diminish attenuation. If monitoring data
and model projections support adequate capacity for 3H
attenuation within the plume, then the site characterization
effort can progress to Tier IV.
Tier IV. Finally, under Tier IV a monitoring plan is estab-
lished along with contingency plans in the event of MNA
failure. It is recommended that the monitoring plan be
designed to establish both continued plume stability and
to identify changes in ground-water levels, flow velocity,
or direction that might influence the efficiency of tritium
removal via radioactive decay. In particular, sites at which
residual tritium sources are left in unsaturated zones should
include monitoring points to assess changes in the release
of tritium to the saturated aquifer due to increased surface
infiltration or rises in the ground-water table. Changes in
system hydraulics may serve as monitoring triggers for
potential MNA failure. In this instance, a contingency plan
can be implemented that incorporates engineered strate-
gies to arrest possible plume expansion beyond compliance
boundaries. Possible strategies to prevent plume expan-
sion include the installation of barriers to minimize tritium
migration from source areas and/or ground-water extraction
with surface treatment.
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learning from a case study: Altamont Hills, California,
U.S.A. Environmental Geology 18:185-194 (1991).
Thibault, D.H., M.I. Sheppard, and PA. Smith. A Critical
Compilation and Review of Default Soil Solid/Liquid
Partition Coefficients, Kd, for Use in Environmental
Assessments. AECL-10125. Whiteshell Nuclear
Research Establishment, Pinawa, Manitoba, Canada
(1990).
USEPA. Understanding Variation in Partition Coefficient,
Kd, Values: Volume II - Geochemistry and Available
Kd Values for Selected Inorganic Contaminants,
EPA/402/R-99/004B, Office of Radiation and Indoor
Air, Washington DC (1999).
USEPA. Inventory of Radiological Methodologies for
Sites Contaminated with Radioactive Materials,
EPA/402/R-06/007, Office of Radiation and Indoor Air,
National Air and Radiation Environmental Laboratory,
Washington, DC (2006). http://www.epa.gov/narel/
IRM_Final.pdf
USEPA. Monitored Natural Attenuation of Inorganic
Contaminants in Ground Water - Volume 1:
Technical Basis for Assessment, EPA/600/R-07/139,
Office of Research and Development, Cincinnati,
OH (2007). http://www.epa.gov/ada/'download/
reports/600RO 7139/600RO 7139. pdf
29
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Visser, A., HP. Broers, and M.F.R Bierkens. Dating
degassed groundwater with3H/3He. Water Resources
Research 43:W10434, doi:10.1029/2006WR005847
(2007).
Visser, A., J.D. Schaap, HP. Broers, and M.F.R Bierkens.
Degassing of 3H/3He, CFCs and SF6 by denitrification:
Measurements and two-phase transport simulations.
Journal of Contaminant Hydrology 103:206-218 (2009).
Warwick, RE., I.W. Croudace, and A.G. Howard. Improved
technique for the routine determination of tritiated
water in aqueous samples. Analytica Chimica
Acta 382:225-231 (1999).
30
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Radon
Daniel I. Kaplan and Robert G. Ford
Occurrence and Distribution
Radon is a noble gas derived from the decay of uranium and
thorium radioisotopes in geologic materials. Three common
radioisotopic forms of this element include 222Rn ("radon"),
220Rn ("thoron"), and 219Rn ("actinon"). Due to very short
half-lives, 220Rn (t1/2 = 55.6 sec) and 219Rn (t1/2 = 3.96 sec)
are not commonly detected in ground water. Radon (222Rn)
is produced from radioactive decay of 226Ra and is part of the
238U decay series. Radon in ground water may be derived
from radioactive decay of mobile forms of 226Ra or due to
ejection from immobile aquifer solids as a result of alpha-
recoil from 226Ra (e.g., Hoehn et al., 1992; Skeppstrom
and Olofsson, 2007). Radon is present in unimpacted
ground-water systems due to the presence of naturally-
occurring 238U in rocks and soils (e.g., Hair and Baldwin,
1995; Veeger and Ruderman, 1998; Hughes et al., 2005;
Campbell, 2006). Uranium concentrations in geological
materials vary greatly (Fayek and Kyser, 1999; Wiegand,
2001) and have been shown by several researchers to have
a strong correlation to ground-water radon concentrations
(e.g., Vinson et al., 2009). For example, the relationship
between granite bedrock and high radon levels has been
observed (e.g., Sasser and Watson, 1978). Elevated radon
concentrations occur near or within mines, especially ura-
nium mines and/or uranium mill tailings (e.g., Landa and
Gray, 1995; USEPA, 2008), and where radioactive waste
is disposed in the subsurface. All categories of radioactive
waste (high-level, low-level, and transuranic waste) contain
concentrated levels of parent isotopes to 222Rn (e.g., see
Figure 3.1).
Geochemistry and Attenuation Processes
Radioactive Decay
Radon is a radioactive noble gas that is colorless and
odorless. Radon has a half-life of 3.82 days and is formed
through alpha decay of 226Ra (t1/2 = 1600 yr) in the 238U decay
series (Figure 3.1). During decay, 222Rn emits a 5.0 MeV
alpha particle. The decay products of 222Rn have short half
lives and variously emit alpha particles (e.g., 6.0 MeV, 218Po;
7.6 MeV 214Po), beta particles (e.g., 295 keV, 214Pb; 352 keV,
214Bi), and/or gamma radiation (e.g., 295 and 352 keV, 214Pb;
609 keV, 214Bi) before forming stable 206Pb (Figure 3.2). As
discussed later, emitted radiation from 222Rn or its short-
lived daughters provides for multiple approaches to quantify
the concentration/activity of this radionuclide.
Aqueous Chemistry
Radon is classified as a noble gas in the periodic table of
elements and is primarily transported as a dissolved gas in
ground water (USEPA, 1999). Of the noble gases, radon
has the highest reported solubility in water (Villalba et al.,
2005). However, radon will tend to partition into air in soil
pores in the variably saturated zone that is in contact with
the ground-water table. Since 222Rn is produced from the
radioactive decay of 226Ra, the mobility and/or accumula-
tion of the parent isotope in the subsurface will, in part,
influence the distribution of 222Rn throughout an aquifer.
As an example, spatial and temporal variability in radon
concentrations in ground water have been linked to redox-
controlled variations in the precipitation or dissolution of iron
and/or manganese (hydr)oxide containing co-precipitated
parent radium (Gainon et al., 2007; Dulaiova et al., 2008).
Radon in water has been demonstrated to preferentially
partition into organic liquids (e.g., Lewis et al., 1987; Davis
et al., 2003). Thus, the relative concentration of 222Rn
observed in ground-water samples may be influenced by the
presence of water-soluble organic solutions or non-aqueous
phase liquids (NAPLs) that may be present within a plume.
As an example, depletions in soil-gas radon concentrations
are used as a diagnostic indicator of the presence of subsur-
face NAPLs (e.g., Schubert et al., 2007; Garcia-Gonzalez
et al., 2008). This variability in 222Rn partitioning behavior
may influence the dimensions of the plume, as well as the
rate of radon transport through the aquifer.
Site Characterization
Overview
Attenuation of radon might be achieved through radioac-
tive decay (Table 3.1). Two factors that will dictate the
adequacy of attenuation via radioactive decay include the
rate of water transport and the total mass and release rate
of radon into the subsurface plume. For some sites, the
volatile transfer of 222Rn from shallow ground water into
overlying unsaturated soil or directly into open or confined
atmospheres may also reduce ground-water concentra-
tions. However, reliance on this mass transfer process will
necessitate determination of the potential consequence of
this exposure pathway to human and/or ecosystem health.
Unacceptable exposures to airborne radon via volatile loss
from ground water may necessitate active control of 222Rn
mass transport through the aquifer. Evaluation of whether
radioactive decay is sufficient to achieve cleanup goals will
31
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Figure 3.1 Potential radioactive decay paths that lead to the production of 222Rn based on data from ICRP (1983).
Radioactive decay in these decay paths occurs via ejection of an alpha (a) or beta particle (p~), or elec-
tron capture (s). The half-life is shown directly below the isotope that is subject to decay; My = megayears,
Ky = kiloyears, y = years, d = days, m = minutes; numbers in parentheses indicate branching fraction.
222Rn
3.823S d
1
a
218Po
3.05m
a
X
214pb
26.8m
X
214Bj
19.9m
X
P~
214Ro
1.64x10-«s
f
ft-
210
a
s
Pb
22.3 y
/
210Bj
5.01 2 d
^
ft-
210Po
1 38.38 d
x
p~
206
a
Pb
stable
Figure 3.2 Daughter radioisotopes resulting from the radioactive decay of 222Rn based on data from ICRP (1983).
Radioactive decay in this series occurs via ejection of an alpha (a) or beta particle (p~). The half-life is
shown directly below the isotope that is subject to decay; y = years, d = days, m = minutes, s = seconds,
stable = non-radioactive isotope.
32
-------
necessitate developing knowledge of the characteristics
of ground-water flow throughout the plume, as well as the
total activity/mass of radon within the plume and entering
from uncontrolled source areas.
Table 3.1 Natural attenuation and mobilization path-
ways for radon.
Attenuation
Processes
Radioactive
decay
Mobilization
Processes
Not applicable
Characterization
Approach
Determination of
ground-water velocity
along relevant
transport pathways
and contaminant
mass release rate
from source areas.
Aqueous Measurements
Relative to determination of the spatial distribution and
temporal variations in radon concentration/activity within
the plume, common methods for the preservation and
analysis of radon in ground water have been reviewed by
Belloni et al. (1995). Three common analysis methods for
direct and indirect determination of 222Rn include 1) liquid
scintillation counting, 2) degassification followed by Lucas
cell counting, and 3) gamma counting. The most commonly
employed method involves the partitioning of 222Rn from the
ground-water sample into either water soluble scintillators
or organic scintillation liquids after radon extraction into
an organic solvent, with counting of decay emissions from
radon and/or its radioactive daughters (e.g., Freyer et al.,
1997; Guiseppe et al., 2006; Leaney and Herczeg, 2006).
For all of these methods, the accuracy of measured 222Rn
concentration/activity will depend on the use of ground-
water sampling and processing methods that result in col-
lection of a well sample that is representative of the aquifer,
and that limit volatile loss of radon gas (e.g., Harris et al.,
2006; Han et al., 2007; Talha et al., 2008).
Several studies have demonstrated measurement of radon
in the field following isolation of 222Rn from ground water
into a gas phase, with detection achieved using a radon-
in-air detector (e.g., Lee and Kim, 2006; Schubert et al.,
2006; Kiliari and Pashalidis, 2008; Schmidt et al., 2008).
The measurement is made using a radon-in-air detector
and conversion of the measured partial pressure of 222Rn
in the gas phase to the corresponding amount of gas dis-
solved in the sampled ground water using Henry's Law.
Important considerations for the reliability of this approach
include accurate measurement or control of temperature
and knowledge that the ground-water matrix can be rea-
sonably represented as pure water. Schubert et al. (2008)
have recently developed an approach by which 222Rn is
sampled directly into the gas phase via diffusion across a
hydrophobic membrane. For low radon concentrations in
surface water, Schubert et al. (2008) found that a porous
membrane tubing length of at least 1-meter would be
adequate for reliable 222Rn quantification. This application
indicates that passive, in-situ sampling of 222Rn within the
well screen may present an alternative to pumping ground
water to the surface.
Partitioning of 222Rn into non-aqueous phase liquids, which
may be a component of the plume or intercepted by the
222Rn plume at down gradient contaminant source areas,
may interfere with analytical methodologies and/or confound
interpretation of plume dimensions. The analytical methods
employed for the detection of radon (directly or via daugh-
ter products) all involve transfer of the radionuclide from
the sample matrix into either a solvent phase or gaseous
matrix. The efficiency of the process to isolate radon will
be influenced by the characteristics of the ground-water
matrix. For sites where transported 222Rn may encounter
non-aqueous phase liquids, it is recommended that sample
collection and analysis procedures be evaluated to ascer-
tain if variations in ground-water matrix could influence the
quality of analytical determinations of 222Rn concentration/
activity throughout the area of investigation.
Solid Phase Measurements
While partitioning of 222Rn to aquifer solids is not known
to occur (excluding free phase NAPLs trapped in the
aquifer matrix), it may be necessary to assess the relative
contributions of this radionuclide from the aquifer matrix
versus dissolved 226Ra or 222Rn released from concentrated
source areas, e.g., infiltration from uranium mill tailings.
The concentration of 222Rn supported by decay of naturally
occurring 226Ra within the aquifer may be ascertained by
ground-water measurements in unimpacted portions of the
aquifer with similar mineralogy and lithology as within the
plume. In addition, aquifer solids may be collected from
within the plume and allowed to equilibrate in a continuous
flow configuration using sampled ground water in which
dissolved 222Rn has been allowed to decay away and dis-
solved 226Ra has been chemically removed. Subsequent
measurements of 222Rn in the column outlet would reflect
the quantity of 222Rn supported by decay and release from
226Ra associated with the aquifer solids, e.g., from natural
and/or anthropogenic forms accumulated via historical
attenuation of 238U and/or 226Ra transported from contami-
nant source areas. This level of characterization may be
useful in ascertaining whether controls on the release of
222Rn or 226Ra from near-surface contaminant zones may
be necessary to allow reliance on radioactive decay for
attenuation of 222Rn in the down gradient plume.
Long-term Stability and Capacity
Since immobilization is not an active mechanism for radon
attenuation in ground water, assessment of long-term
stability is not a factor. The long-term capacity for radon
attenuation within a plume will be dictated by the relative
rate of ground-water flow along relevant transport pathways
compared to the rate for radioactive decay, given a known
flux of radon entering the plume. Thus, a critical factor for
assessing the overall capacity of the aquifer for attenuation
33
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will be evaluation of the mass flux of radon moving through
the plume relative to the rate of water movement through
the aquifer. Radioactive decay may be sufficient to prevent
plume expansion, but this is not likely for sites with an
uncontrolled source of radon entering the subsurface and/
or characteristic times for ground-water transport that are
significantly shorter than the half-life of radioactive decay for
222Rn. Relative to radon release from uncontrolled source
areas, knowledge of the total contaminant mass as well as
the rate and frequency of release into the saturated aquifer
needs to be developed. In addition, translocation of 226Ra
through the aquifer due to changes in chemistry may influ-
ence the dimensions and location of the observed 222Rn
plume. In order to make a reliable assessment of the mass/
activity flux of radon into the plume, it will be important to
understand the characteristics of the hydrogeologic system
and the dynamics of water and contaminant transfer from
contaminant source areas into the plume.
Tiered Analysis
Determination of the viability of radon remediation in ground
water via monitored natural attenuation will depend upon
proper assessment of contaminant loading to the aquifer
relative to the velocity of ground-water travel along relevant
transport pathways to the point(s) of compliance. The
goal of site assessment is to demonstrate that radioactive
decay is adequate to meet cleanup goals given current and
projected hydrologic conditions for the site. The following
tiered analysis structure for site characterization provides
an approach to evaluate candidate sites and define the
potential limitations of MNA as part of a remedy for ground-
water cleanup.
Tier I. Site characterization under Tier I will involve dem-
onstration that the ground-water plume is static or shrink-
ing, has not reached compliance boundaries, and does
not impact existing water supplies. It is also important
at this stage of the site evaluation to determine source
term characteristics such as the inventory of contaminant
mass and the current and historical rate of release into the
subsurface. Acquisition of this information in combination
with identification of a stable plume provides justification
for proceeding to Tier II characterization efforts.
Tier II. Under Tier II, the apparent rate of attenuation is
determined. Estimates of a site attenuation rate(s) can be
assessed via a well transect along the ground-water flow
path. In addition, time-series data may be collected at
one or more monitoring points within the plume (USEPA,
2007; Section IIIA.5). Since radioactive decay will be the
dominant mechanism controlling radon attenuation, deter-
mination of the velocity of ground-water flow along relevant
transport pathways will be critical to evaluating the potential
cleanup time frame. This information will allow assessment
of the relative timescales for contaminant attenuation and
fluid transport and determination of whether remediation
objectives can be met within the required regulatory time
frame. Determination of the adequacy of radioactive
decay to achieve cleanup goals will necessitate detailed
analysis of system hydrology relative to flow pathway(s),
flow velocity, and temporal variations in flow velocity and/or
direction within the boundaries of the plume. This informa-
tion, in combination with knowledge of contaminant source
location(s), radon mass, and emanation characteristics
from source zones, can be employed to develop a decay-
transport model to project radon activity/concentration
distribution throughout the plume. The demonstration of
concurrence between conceptual and mathematical mod-
els describing radon transport will entail development of
site-specific parameterization of ground-water flow along
relevant transport pathways.
Tier III. Under Tier III, it will be important to assess whether
the capacity of the system is adequate to sustain radon
attenuation (e.g., prevent plume expansion) relative to the
mass of the contaminant being transported through the
plume. Two principal factors that may influence capac-
ity include changes in water transport and/or changes in
radon mass/activity flux entering the plume. Source release
characteristics may be influenced by the physical location
of sources within the subsurface relative to the ground-
water table. Fluctuations in infiltration through shallow,
unsaturated zones and/or water table elevations within the
aquifer due to variations in recharge may lead to periodic
increases in radon flux into ground water from contaminant
source areas. Changes in land usage (including at recharge
zones, wetlands, and over the contaminated site) and/or
ground-water withdrawals from the aquifer may influence
ground-water flow direction and velocity (USEPA, 2007;
Section IMA), which in turn directly influences radon travel
time. It is recommended that additional radon transport
modeling be included to evaluate the impact of these vari-
ous scenarios to be assured that these perturbations do
not significantly diminish attenuation. If monitoring data
and model projections support adequate capacity for radon
attenuation within the plume, then the site characterization
effort can progress to Tier IV.
Tier IV. Finally, under Tier IV a monitoring plan is estab-
lished along with contingency plans in the event of MNA
failure. It is recommended that the monitoring plan be
designed to establish both continued plume stability and
to identify changes in ground-water levels, flow velocity,
or direction that might influence the efficiency of radon
removal via radioactive decay. In particular, sites at which
residual, subsurface 226Ra sources are left near the water
table should include monitoring points to assess changes
in the release of radon to the saturated aquifer due to
increased surface infiltration or rises in the ground-water
table. Changes in system hydraulics may serve as moni-
toring triggers for potential MNA failure. In this instance,
a contingency plan can be implemented that incorporates
engineered strategies to arrest possible plume expansion
beyond compliance boundaries. Possible strategies to
prevent plume expansion include the removal of 226Ra from
periodically saturated source areas and/or ground-water
extraction with surface treatment.
34
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P. Santaroni, G. Torri, and R. Vasselli. Optimization
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36
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Strontium
Patrick V. Brady, James E. Amonette, Robert G. Ford, Richard T. Wilkin
Occurrence and Distribution
90Sr is produced in nuclear reactors as a product from
fission of 235U at a yield of about 6% (or 6 atoms of 90Sr
for every 100 fissions of 235U). It is present in wastes and
wastewaters associated with fuel handling and reprocess-
ing. Otherwise, the primary inputs of 90Sr to the environ-
ment are from nuclear bomb fallout, nuclear accidents
(e.g., Chernobyl), and releases from nuclear power plant
and nuclear propulsion emissions. Average 90Sr activity
in soils is 100 mCi/m2 (USEPA, 1999). Non-radioactive
(or stable) isotopes of Sr include 88Sr (82.74% of total Sr),
86Sr (9.75%), 87Sr (6.96%), and 84Sr (0.55%). 90Sr is often
a primary radioisotope of concern at low-level radioactive
waste storage facilities because of its radiotoxicity, its rela-
tively high mobility, and its general abundance in low-level
wastes. Like stable isotopes of strontium, 90Sr is always
present in ground water in a divalent oxidation state [Sr(ll)].
It forms weak aqueous complexes with carbonate, sulfate,
and chloride and is typically present in natural waters as
an uncomplexed cation. Because of similar charge and
radius, strontium tends to mimic calcium in the environment.
Consequently 90Sr is often observed as an exchangeable
cation on clay minerals, as a component of carbonate or
sulfate minerals, and to a much lesser extent, sorbed weakly
to iron (hydr)oxides and other minerals.
Geochemistry and Attenuation Processes
Radioactive Decay
Two radioactive isotopes of strontium that result from fis-
sion reactions during nuclear materials production are 89Sr
(50.5 day half-life; decays to stable 89Y) and 90Sr (28.8 year
half-life). Of these radioisotopes, 90Sr has been observed
in ground water due to its longer half life, decaying to 90Y
and emitting a 546 keV maximum energy beta particle in
the process. Subsequently, 90Y decays with a half-life of 64
hours to stable 90Zr (Figure 4.1). Several stable isotopes
of strontium exist in nature (USEPA, 1999). The most
prevalent stable isotope is 88Sr, comprising about 82.6%
of natural strontium. The other three stable isotopes and
their relative abundance are 84Sr (0.6%), 86Sr (9.9%), and
87Sr (7.0%). The majority of total dissolved strontium in
ground water will typically consist of the stable isotopes.
As an example, stable strontium concentrations in ground
water for aquifers not impacted by radioactive wastes have
been reported with ranges of 110-6120 |ag/L (Jacobson
and Wasserburg, 2005) and 14-43,500 |ag/L (Mclntosh
and Walter, 2006). This compares to the mass-equivalent
risk based concentration limit for 90Sr of 0.000000059 |ag/L
(see Table 1.1).
90S|.
29.1 y
64.0 h
90Zr
stable
Figure 4.1 Decay series for 90Sr based on data from
ICRP (1983). Radioactive decay in this series
occurs via ejection of a beta particle (p~). The
half-life is shown directly below the isotope
that is subject to decay; y = years, h = hours,
stable = non-radioactive isotope.
Aqueous Speciation
Strontium predominantly occurs as an uncomplexed cat-
ion in solution (USEPA, 1999; Siegel and Bryan, 2003).
There is some tendency for strontium to form carbonate
complexes, but this is predicted to occur at relatively high
pH (Figure 4.2; Felmy et al., 1998). Strontium can form
complexes with organic constituents in solution, such as
organic acids (acetate; Ragnarsdottir et al., 2001) or syn-
thetic chelating agents (EDTA; Felmy and Mason, 2003).
However, these complexes are not likely to play a major
role in ground water, where more stable complexes of EDTA
with other cations in solution would out-compete strontium
complexation. As demonstrated by Pace et al. (2007), even
strong chelating agents like EDTA appear to exert only a
short-term influence on strontium aqueous speciation due
to competition with other cations in solution that form more
stable complexes with EDTA. There is indirect evidence
that 90Sr may bind to natural organic compounds such as
fulvic acid (Zhao and Chen, 2006), but the overall influence
of these compounds on the aqueous speciation of stron-
tium will likely be limited to near-surface systems with high
concentrations of dissolved organic carbon.
37
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Solubility
Strontium may precipitate with dissolved carbonate or sul-
fate to form the minerals strontianite (SrCO3) and celestite
(SrSO4), respectively. However, this is only anticipated for
ground water with elevated concentrations of strontium and
carbonate/sulfate. As shown in Figure 4.2a, precipitation
of strontianite generally occurs under alkaline pH in ground
water where natural stable strontium concentrations are
elevated. In contrast, strontium may be coprecipitated
during formation of carbonates of calcium or ferrous iron,
e.g., under conditions of elevated alkalinity and/or ferrous
iron that develop during microbial degradation of organic
compounds (e.g., Roden et al., 2002; Fujita et al., 2004). As
shown in Figure 4.2b, the precipitation of siderite (FeCO3) or
calcite (CaCO3) would likely occur prior to the precipitation
of strontianite from ground water. The degree that strontium
may partition into these carbonates will be governed by
competition with other divalent cations in solution that may
be present at higher concentrations (Roden et al., 2002),
although coprecipitation with calcite has been observed in
natural systems (Ferris et al., 1995).
Adsorption
Because of the relatively high solubility of strontium-bear-
ing minerals, adsorption tends to play a more important
role in limiting strontium concentration in ground water.
Cation exchange onto clay minerals is often the dominant
adsorption reaction during transport through an aquifer
(e.g., Cerling and Spalding, 1982; Serne and LeGore,
1996; USEPA, 1999; McKinley et al., 2007). Laboratory
studies suggest that strontium may bind to the surface of
oxide minerals (e.g., Carroll et al., 2008) and evaluation
of biogenic iron (hydr)oxides collected from subsurface
systems support the potential for this adsorption reaction
under certain conditions (e.g., Ferris et al., 2000). Solid
phase characterization data from some field studies of 90Sr
plumes suggest that strontium adsorption reactions may
involve a combination of cation exchange onto clay miner-
als in combination with surface complexation reactions on
oxyhydroxide minerals (e.g., Jackson and Inch, 1989).
-4
(a)
-5
CO
cn
o
-7
-8
Sr concentration modeled in (b)
Sr2*
Strontianite
\"\ 1
SrCO3(aq)
25°C
(b)
Figure 4.2 (a) Solubility and speciation of strontium
as a function of pH and PCO2 (strontian-
ite, SrCO3). (b) Strontianite stability field
(Sr2+=10-em, 0.087 mg L1) in relation to sta-
bility fields for calcite (CaCO3, variable Ca2+
concentration) and siderite (FeCO3, variable
Fe2* concentration, FeO suppressed). Diagram
constructed using the LLNL thermodynamic
database (therm.com.V8.R6+) modified using
the recommendations of Tournassat et al.
(2008) for the solubility of strontianite.
38
-------
Site Characterization
Overview
Attenuation of 90Sr might be achieved through radioactive
decay with the influence of adsorption or coprecipita-
tion reactions. In general, since strontium adsorption is
dominated by highly reversible ion exchange reactions,
it is anticipated that the primary function of this mecha-
nism would be to limit the rate of mass transport to allow
radioactive decay to remove sufficient contaminant mass.
Co-precipitation reactions are likely to occur only under situ-
ations of high alkalinity production in concert with elevated
concentrations of constituents such as calcium or ferrous
iron. These processes would typically only be relevant to
plumes with elevated microbial activity due to degradation
of organic constituents also transported within the plume.
A list of potential attenuation processes is provided in
Table 4.1. Two factors that will dictate the adequacy of
attenuation via radioactive decay and/or co-precipitation
include the rate of water transport and the total mass and
release rate of 90Sr into the subsurface plume.
Table 4.1 Natural attenuation and mobilization path-
ways for strontium.
Attenuation
Processes
Radioactive
decay
Adsorption
onto aquifer
minerals (clay
minerals, iron
oxy hydroxides)
Co-precipitation
with Fe(ll)
carbonate
or calcium
carbonate
Mobilization
Processes
Not
applicable
Desorption
due to
decreasing
pH or
competition
from major
cations
in ground
water.
Dissolution
due to
decreasing
pH or
oxidation
of Fe(ll)
carbonate
Characterization
Approach
Determination
of ground-water
velocity along
relevant transport
pathways and
contaminant mass
release rate from
source areas.
Evaluate total
adsorption capacity
of aquifer solids
under representative
ground-water
chemistry; chemical
extractions
to assess
concentrations of
exchangeable 90Sr
in aquifer solids
along relevant
transport pathways.
Evaluate formation
of carbonate
minerals along
relevant transport
pathways and
determine 90Sr
association with this
mineral fraction.
Aqueous Measurements
As a beta particle emitter (mean and end-point energy of
195.8 keV and 546.0 keV, respectively), 90Sr can be quan-
tified using radiometric analysis (USEPA, 2004; USEPA,
2006). However, the presence of other beta-emitting
radionuclides in the sample can interfere with the determi-
nation of the activity/concentration of 90Sr via detection of
beta emissions due to overlap in their corresponding beta-
emission spectra (Horwitz et al., 1991; Tinker et al., 1997).
Examples of other beta-emitting radionuclides that may be
present in the original sample include 90Y (decay daughter;
232.5 keV mean, 643.2 keV end-point) or products from
natural/anthropogenic uranium-thorium decay series (e.g.,
214Pb, 223 keV mean, 26.8 m half-life; 214Bi, 642 keV mean,
19.9 m half-life). Procedures exist for the isolation of 90Sr
from liquid sample matrices (USEPA, 2004; USEPA, 2006),
but in-growth of 90Y following separation and the potential
for incomplete isolation of Pb radioisotopes may need to
be addressed (Beals et al., 2001). Sample storage to allow
in-growth of 90Y to secular equilibrium with 90Sr (typically
several weeks with detection of 90Y) can address potential
bias from 90Y that may not initially be in secular equilibrium
with 90Sr (e.g., Fiskum et al., 2000). This storage time
may also address the short-lived radioisotopes from the
uranium-thorium decay series (e.g., 214Pb and 214Bi), but it
may not necessarily address interference from 210Pb (e.g.,
Goutelard et al., 2000). In the absence of interfering beta
emitters, there are reported methods for on-line separation
and simultaneous detection of 90Sr and 90Y via chromato-
graphic techniques with simultaneous measurement of both
radioisotopes (e.g., Egorov et al., 2006; Grate et al., 2008).
Solid phase extraction resins may also be employed to
isolate 90Sr with immediate analysis of reacted resin (e.g.
Fiskum et al., 2000; Beals et al., 2001), but the adsorption
of non-targeted beta-emitting radionuclides needs to be
initially evaluated for the various water chemistries that
might be analyzed (e.g., Beals et al., 2001; Egorov et al.,
2006). For solid phase extraction or other approaches to
isolate 90Sr from the water sample, a known trace mass of
radioactive 85Sr is commonly spiked into the sample matrix
in order to assess chemical recovery of 90Sr during sample
processing.
Recent developments in the ICP-MS methodology allow
detection of 90Sr at levels comparable to reporting limits
for radiometric techniques (Taylor et al., 2007; Hou and
Roos, 2008). The improvement in detection limit has been
achieved through the use of procedures for off-line and on-
line concentration of 90Sr in the analyzed sample (similar to
methods employed for radiometric techniques) and through
use of dynamic reaction cell configurations to control forma-
tion of molecular ions of similar mass that form in the plasma
prior to introduction to the mass spectrometer. For example,
Taylor et al. (2007) have demonstrated use of oxygen intro-
duced into the plasma to convert 90Zr (mass 90) to 90ZrO
molecular ions (mass 106), eliminating the positive bias to
measured 90Sr mass from 90Zr that may be present in the
sample. The presence of 90Y does not significantly interfere
with this approach given the low mass abundance of 90Y at
39
-------
secular equilibrium (i.e., 90Y/90Sr mass ratio will be 0.025%
at secular equilibrium). An additional benefit to using mass
spectrometry is the ability to simultaneously monitor other
stable and radioactive isotopes in the matrix. This analysis
facilitates evaluation of the chemical conditions at locations
where subsurface samples were retrieved, as well as make
use of stable and/or radioactive isotope ratios as signatures
of sources of water throughout the plume (e.g., Singleton
et al., 2006; Christensen et al., 2007).
Solid Phase Measurements
Solid phase measurements that may provide information
useful to assessing processes controlling 90Sr retardation
within the aquifer and the capacity along relevant transport
pathways include the determination of 90Sr partitioning to
aquifer solids, the cation-exchange capacity of aquifer
solids, and identification of aquifer solids mineralogy that
may participate in adsorption and/or precipitation reac-
tions. Evaluation of the mass distribution of 90Sr between
co-located ground water and aquifer solids throughout the
plume provides an assessment of the extent that retarda-
tion reactions limit strontium migration. Bulk solid-phase
partitioning can be conducted using total digestion or acid-
extractable techniques (e.g., Friberg, 1997; Solecki, 2007)
with appropriate measures to assess extraction efficiency
and potential interference from other beta-emitters that
may interfere with radiometric techniques (Chang et al.,
2004; USEPA, 2004; USEPA, 2006). For these approaches,
interference from matrix constituents, including other beta-
emitters, could be significant due to release of these con-
stituents as a result of dissolution of the solid matrix. As with
aqueous measurements, these potential interferences need
to be considered in the design of the analytical procedure.
There are a range of aqueous reagent solutions that may
be employed to measure the total cation exchange capacity
of the aquifer solids, as well as the fraction of exchange-
able 90Sr (e.g., Cerling and Spalding, 1982; McKinley et
al., 2007). An assumption of these procedures is that only
exchangeable cations are released from the solid matrix.
It is recommended that the potential for partial dissolution
of the more labile mineral fraction in aquifer solids during
extraction (e.g., Jackson and Inch, 1989) be assessed.
This may be achieved through measurement of the major
ion chemistry in the extract solution (e.g., iron, manganese,
sulfate), which may serve as markers for the dissolution of
oxides, sulfides or other components that may bind 90Sr in
a less labile form. Physical and chemical procedures, e.g.,
size f ractionation or selective extraction of carbonate miner-
als, may also be employed to assist in identifying specific
components with aquifer solids that dominate strontium
solid-phase partitioning. Examples of procedures to identify
the type and abundance of specific minerals along transport
pathways are available in the literature (e.g., Cerling and
Spalding, 1982; Jackson and Inch, 1983; McKinley et al.,
2007). Additional information on analysis approaches and
analytical techniques applied to solid phase characterization
is provided in USEPA (2007; Section NIB).
Long-term Stability and Capacity
The long-term stability of 90Sr attenuated through adsorption
or coprecipitation will depend upon the stability of the host
mineral and the abundance of other ions which might dis-
place adsorbed strontium. The most easily envisioned case
of the first is that 90Sr coprecipitated in carbonates might
be remobilized if ambient pH were to decrease. Increases
in divalent cation levels can be expected to work against
90Sr immobilization for sites dominated by cation exchange
reactions. Reductive dissolution of iron (hydr)oxides might
cause remobilization of adsorbed 90Sr. Review of the extent
of plume development for a number of sites indicates that
cation exchange within the saturated aquifer may have
insufficient stability to prevent plume expansion (Brady et
al., 2002). Thus, a critical factor for assessing the overall
capacity of the aquifer for attenuation will be evaluation of
the mass flux of 90Sr moving through the plume relative to
the rate of water movement through the aquifer. Radioactive
decay may be sufficient to prevent plume expansion, but
this is not likely for sites with an uncontrolled source of
90Sr entering the subsurface and/or characteristic times for
ground-water transport that are significantly shorter than
the half-life of radioactive decay.
Tiered Analysis
Determination of the viability of 90Sr remediation in ground
water via monitored natural attenuation will depend upon
proper assessment of contaminant loading to the aquifer
relative to the velocity of ground-water travel and the prevail-
ing geochemistry and mineralogy within the contaminant
plume and the down gradient zone prior to the point(s) of
compliance. The goal of site assessment is to demonstrate
the process(es) controlling strontium sequestration onto
aquifer solids and the long-term stability of solid phase
strontium as a function of existing and anticipated ground-
water chemistry. The following tiered analysis structure
for site characterization provides an approach to evaluate
candidate sites and define the potential limitations of MNA
as part of a remedy for ground-water cleanup.
Tier I. Site characterization under Tier I will involve demon-
stration that the ground-water plume is static or shrinking,
has not reached compliance boundaries, and does not
impact existing water supplies. Once this is established
through ground-water characterization, evidence is col-
lected to demonstrate 90Sr partitioning to aquifer solids
within the plume. If immobilization processes are active
throughout the plume, then there should be an observed
increase in solid phase concentrations within regions of the
plume with higher aqueous concentrations, e.g., near the
source term. Evaluation of the mass/activity of 90Sr distrib-
uted between ground water and aquifer solids throughout
the plume is recommended to account for both existing
and potentially mobile forms of 90Sr. This field partitioning
data may be supplemented by geochemical modeling that
incorporates measured water chemistry (e.g., pH and major
ion chemistry) throughout the plume to assess the potential
for solubility control by precipitation of carbonate minerals.
It is also important at this stage of the site evaluation to
40
-------
determine source term characteristics such as the inventory
of contaminant mass and the current and historical rate of
release into the subsurface. Acquisition of this information
in combination with identification of a stable plume provides
justification for proceeding to Tier II characterization efforts.
Tier II. Under Tier II, the apparent rate and mechanism(s)
of attenuation are determined. Estimates of a site attenu-
ation rate(s) can be assessed via a well transect along the
ground-water flow path. In addition, time-series data may
be collected at one or more monitoring points within the
plume (USEPA, 2007; Section IIIA.5). This information will
allow assessment of the relative timescales for contaminant
attenuation and fluid transport and determination of whether
remediation objectives can be met within the required
regulatory time frame. As part of this analysis, it is critical
to determine the relative contributions of radioactive decay
and immobilization to the overall observed attenuation.
Determination of the contribution of radioactive decay will
necessitate detailed analysis of system hydrology relative
to flow pathway(s), flow velocity, and temporal variations in
flow velocity and/or direction within the boundaries of the
plume. This information, in combination with knowledge
of contaminant source release characteristics, can be
employed to develop a decay-transport model to project 90Sr
activity/concentration distribution throughout the plume in
the absence of adsorption/coprecipitation processes. For
systems in which immobilization plays a role in observed
attenuation, it will be necessary to identify whether adsorp-
tion onto existing aquifer minerals or coprecipitation with
newly formed minerals predominates. This effort will require
determination of the chemical speciation of solid phase 90Sr
and may be approached according to the following scheme:
1)
2)
Calculation of saturation state of ground water
relative to precipitation of carbonate or (hydr)oxide
minerals along relevant 90Sr transport pathways;
Determination of aquifer mineralogy to determine
the relative abundance of components with
documented capacity for strontium adsorption (e.g.,
Jackson and Inch, 1983; 1989), with implementation
of steps for aquifer solids collection, processing
and analysis that avoid transformation of mineral
species from reduced zones (e.g., oxidation of
ferrous carbonate to ferric (hydr)oxide); and
3)
Determination of 90Sr-sediment associations via
chemical extractions designed to target specific
components within the aquifer sediment (e.g.,
Cerling and Spalding, 1982; McKinley et al., 2007).
This compilation of information will facilitate identification of
the reaction(s) leading to 90Sr attenuation. The demonstra-
tion of concurrence between conceptual and mathematical
models describing strontium transport (both stabile and
radioactive isotopes) will entail development of site-specific
parameterization of the chemical processes controlling 90Sr
solid phase partitioning.
Tier III. Once the contributions from radioactive decay
and adsorption/coprecipitation processes have been deter-
mined, the subsequent characterization effort under Tier III
will involve determination of the stability of immobilized 90Sr
and the capacity of the aquifer to sustain continued uptake.
It is recommended that the stability of immobilized 90Sr be
tested based on the anticipated evolution of ground-water
chemistry concurrent with plume evolution. For example,
changes in ground-water pH and/or cation composition can
exert a significant influence on 90Sr adsorption. Therefore,
it is recommended that sediment leach tests be conducted
to characterize the magnitude of 90Sr re-mobilization as a
function of pH for a ground-water chemistry representative
of site conditions (including stable strontium isotopes pres-
ent in ground water). It is recommended that the capacity
for 90Sr uptake onto aquifer solids be determined relative to
the specific mechanism(s) identified in Tier II. For sites in
which a continuing source of 90Sr to the saturated aquifer
exists, it is recommended that potential steps to minimize
or eliminate this continued contaminant flux be evaluated
and implemented where feasible. If site-specific tests
demonstrate that the stability and capacity for 90Sr immo-
bilization, in combination with continued elimination of 90Sr
via radioactive decay, are sufficient to sustain attenuation,
then the site characterization effort can progress to Tier IV.
Tier IV. Finally, under Tier IV a monitoring plan is estab-
lished along with contingency plans in the event of MNA
failure. It is recommended that the monitoring plan be
designed to establish both continued plume stability and to
identify changes in ground-water chemistry that may lead to
re-mobilization of attenuated strontium (both 90Srand stable
strontium isotopes). The specific chemical parameters to be
monitored will include those identified under Tier III that may
halt strontium partitioning to aquifer minerals and/or result
in solubilization of precipitates into which strontium has
been incorporated. Solution phase parameters that could
alter either strontium coprecipitation or adsorption include
changes in pH or concentrations of competing cations in
ground water. As an example, increases in calcium concen-
trations in ground water could signal either 1) the potential
for displacement of 90Sr from cation exchange sites, or
2) the dissolution of calcium carbonate minerals in which
strontium is coprecipitated. Changes in water chemistry
may occur prior to observed changes in solution 90Sr and,
thus, serve as monitoring triggers for potential MNA failure.
In this instance, a contingency plan can be implemented
that incorporates engineered strategies to arrest possible
plume expansion beyond compliance boundaries. Possible
strategies to prevent plume expansion include ground-water
extraction with surface treatment, installation of permeable
reactive barriers to enhance uptake capacity perpendicu-
lar to the direction of plume advance, or enhancement of
coprecipitation processes within the aquifer through the
injection of soluble reactive components.
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Serne, R.J. and V.L. LeGore. Strontium-90 Adsorption-
Desorption Properties and Sediment Characterization
at the 100 N-Area. PNL-10899, Pacific Northwest
Laboratory, Richland.WA (1996). http://www.osti.gov/
bridge/servlets/purl/186728-XT8Xqv/webviewable/
Siegel, M.D. and C.R. Bryan. Environmental Geochemistry
of Radioactive Contamination. SAND2003-2063,
Sandia National Laboratories, Albuquerque, NM (2003).
http://www.prod. sandia. gov/cgi-bin/techlib/access-
control.pl/2003/032063.pdf
Singleton, M.J., K. Maher, D.J. DePaolo, M.E. Conrad,
and PE. Dresel. Dissolution rates and vadose zone
drainage from strontium isotope measurements of
groundwater in the Pasco Basin, WA unconfined aquifer.
Journal of Hydrology 321:39-58 (2006).
Solecki, J. Studies of 90Sr concentration and migration in
the soils of the Leczna-Wlodawa Lake District. Journal
of Radioanalytical and Nuclear Chemistry 274:27-38
(2007).
Taylor, V.F, R.D. Evans, and R.J. Cornett. Determination
of 90Sr in contaminated environmental samples by
tuneable bandpass dynamic reaction cell ICP-MS.
Analytical and Bioanalytical Chemistry 387:343-350
(2007).
Tinker, R.A., J.D. Smith, and M.B. Cooper. Determination
of strontium-90 in environmental samples containing
thorium. Analyst 122:1313-1317 (1997).
Tournassat, C., C. Lerouge, P. Blanc, J. Brendle, J.-M.
Greneche, S. Touzelet, and E.G. Gaucher. Cation
exchanged Fe(ll) and Sr compared to other divalent
cations (Ca,Mg) in the bure Callovian-Oxfordian
formation: Implications for porewater composition
modelling. Applied Geochemistry 23:641-654 (2008).
USEPA. Understanding Variation in Partition Coefficient,
Kd, Values: Volume II - Geochemistry and Available
Kd Values for Selected Inorganic Contaminants,
EPA/402/R-99/004B, Office of Radiation and Indoor Air,
Washington DC (1999). http://www.epa.gov/rpdwebOO/
deanup/402-r-99-004. html
USEPA. Multi-Agency Radiological Laboratory Analytical
Protocols Manual, Volume II: Chapters 10-17 and
Appendix F, EPA/402/B-04/001B (NUREG-1576),
Washington, DC (2004). http://www.epa.gov/radiation/
marlap/manual. html
USEPA. Inventory of Radiological Methodologies for
Sites Contaminated with Radioactive Materials,
EPA/402/R-06/007, Office of Radiation and Indoor Air,
National Air and Radiation Environmental Laboratory,
Washington, DC (2006). http://www.epa.gov/narel/
IRM_Final.pdf
USEPA. Monitored Natural Attenuation of Inorganic
Contaminants in Ground Water - Volume 1 Technical
Basis for Assessment. EPA/600/R-07/139, U.S.
Environmental Protection Agency, Cincinnati,
OH (2007). http://www.epa.gov/ada/'download/
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Zhao, D. and C. Chen. Effect of fulvic acid on the sorption
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Radioanalytical and Nuclear Chemistry 270:445-452
(2006).
43
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Technetium
Patrick V. Brady, James E. Amonette, Robert G. Ford, Richard T. Wilkin
Occurrence and Distribution
"Tc is a fission product that is produced in nuclear power
plants as a product from thermal neutron fission of 235U
(6.1% yield) and fission of 239Pu (5.9% yield; IAEA, 2006).
The metastable nuclear isomer of "Tc, 99mTc, has been
used for many years in diagnostic nuclear medical imaging
procedures due to emission of a 140 keV gamma ray during
its decay to 99Tc (Patton et al., 1966). 99mTc is a product of
the decay of 99Mo produced in nuclear fission reactions or
it can be produced using a commercial medical cyclotron
(e.g., Beaver and Hupf, 1971). There are no stable, natu-
rally occurring isotopes of technetium. In nuclear reactor
operations, 99Tc is produced through several processes,
and is often found in ion-exchange resins, filter sludges,
evaporator bottoms, cartridge filters, trash, and decom-
missioning wastes. 99Tc is present in a number of the
high-level waste tanks at the Hanford Reservation and is
a constituent of several ground-water plumes at this site
(e.g., Hartman et al., 2006). Because it is long-lived and
rapidly transported under oxidizing conditions, 99Tc tends
to dominate performance assessment calculations that are
used to predict doses that might result from radionuclide
releases from high and low-level waste facilities.
Geochemistry and Attenuation Processes
Radioactive Decay
99Tc has a half-life of 213,000 years and decays to stable
99Ru with the emission of a 293 keV maximum energy
beta particle (ICRP, 1983; Eckerman et al., 1994; USEPA,
2002). The majority of technetium is generated as 99mTc
from decay of the 99Mo fission product. This metastable
form (6.02 h half-life) decays to 99Tc with the emission of a
140 keV gamma-ray Based on a Maximum Contaminant
Level (MCL) of 4 millirem per year for beta particle and
photon radioactivity from man-made radionuclides in drink-
ing water, one can estimate an equivalent activity-based
MCL of 900 pCi/L for 99Tc (mass equivalent concentration
of 0.053 |j.g/L) assuming 99Tc as the only beta- or photon-
emitting radioisotope (USEPA, 2002).
Aqueous Speciation
99Tc can exist in eight oxidation states ranging from Tc(-l)
to Tc(VII). The two most common oxidation states are
Tc(VII)andTc(IV) under oxidizing and reducing conditions,
respectively. Reviews of the aqueous geochemistry of tech-
netium have been provided in USEPA (2004a), Siegel and
Bryan (2003), and IAEA (2006). The pertechnetate anion
(TcO4~) is the dominant chemical form of dissolved Tc(VII)
in ground water. This anion has not been observed to form
complexes of consequence in ground-water systems. There
is evidence to support formation of soluble complexes with
inorganic or organic ligands following reduction to Tc(IV).
Formation of soluble Tc(IV) complexes with bicarbonate/
carbonate is possible in ground water with elevated alkalin-
ity (Wildung et al., 2000). Natural and synthetic chelating
agents may also complex with Tc(IV), e.g., humic/fulvic
acids and EDTA/NTA, respectively. Microcosm studies
conducted by Maset et al. (2006) suggest that synthetic
chelating agents would likely have minimal influence on
the chemical speciation of Tc(IV) due to the presence of
competing solid sorbents for aqueous Tc(IV) and/or major
cation competition for complexation with EDTA/NTA. In
contrast, there is evidence that natural dissolved organic
compounds such as humic acid may form relatively stable
complexes with reduced technetium (e.g., Maes et al., 2003;
Gu and Ruan, 2007; Geraedts and Maes, 2008). These
complexes may enhance Tc(IV) transport in organic-rich
ground water under conditions unfavorable to sorption
onto aquifer solids or where exposure to dissolved oxygen
results in oxidation of complexed technetium (Gu and Ruan,
2007). In addition, these complexation reactions will be in
competition with the tendency for Tc(IV) to form hydrous
oxides with low solubility.
Solubility
The formation of precipitates incorporating Tc(VII) are not
likely to form in ground water. In contrast, formation of
hydrous oxides of Tc(IV) have been observed in soils and
sediments influenced by microbial respiration of natural
or anthropogenic sources of organic carbon (Istok et al.,
2004; Abdelouas et al., 2005; Begg et al., 2008; Morris et
al., 2008) or due to abiotic reduction in the presence of
electron donors such as ferrous iron (Lloyd et al., 2000;
Zachara et al., 2007; Peretyazhko et al., 2008). As a point
of reference, the stability field for TcO2»2H2O is shown
in Figure 5.1 a when other Tc(IV) precipitate phases are
suppressed from the speciation calculations. When all
precipitate phases are allowed to form, a range of solids
with varying technetium oxidation state are predicted to form
(Figure 5.1 b). However, many of the mixed- or lower-oxida-
tion state solids for which solubility constants are published
have not been observed to form in ground-water systems
and it is recommended that their predicted presence based
on model calculations be viewed as highly uncertain. The
influence of different alkalinity levels in ground water on the
stability of TcO2»2H2O is shown in Figure 5.2a. Based on
these model projections, the formation of Tc(lV)-carbonate
complexes could limit precipitation of this hydrous oxide
in systems with alkaline pH and high carbonate alkalinity.
45
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Available thermodynamic databases list solubility con-
stants for technetium sulfides (e.g., Guillaumont et al.,
2003; Rard, 2005), although recent critical reviews point
to uncertainty in the reliability of these data (e.g., IAEA,
2006). Several experimental studies conducted in sulfidic
systems indicate the potential formation of "Tc(IV)S2-like"
precipitates based on spectroscopic characterization of the
reaction products (e.g., Wharton et al., 2000; Livens et al.,
2004). These observations are consistent with formation
of "Tc(IV)Sx" in high-level waste systems with a structural
stoichiometry equivalent to Tc(IV)3S(-ll)2(S(-l)2)4, nominally
Tc3S10 (Lukens et al., 2005). It should be noted that Lukens
et al. (2005), based on structural analysis of technetium
sulfide phases in these systems, call into question the
formation of Tc(VII)2S7, which is commonly documented
in available thermodynamic databases. These observa-
tions contrast with identification of Tc(VII) as the oxidation
state in a Tc2S7 reference material (Wharton et al., 2000)
and the recently observed formation of the phase in an
experimental study (Liu et al., 2007). In addition, Liu et
al. (2008) suggest that a Tc(IV)O2-like phase may be the
precipitate that forms upon interaction with Fe(ll) sulfide.
As shown in Figure 5.2b, thermodynamic calculations
indicate that formation of technetium sulfide precipitates
could compete with formation of the hydrous Tc(IV) oxide
under sulfate-reducing conditions. In general, the ability to
project what type of reaction controls precipitation of Tc(lV)
with any certainty is limited for sulfate-reducing systems,
although formation of TcO2»2H2O seems best supported.
Adsorption
Under oxidizing conditions, technetium is soluble and typi-
cally does not adsorb onto aquifer solids. The sorption Kd
for "Tc is typically assumed to be zero (USEPA, 2004a).
Values of Kd>0 that have been reported in the literature
have typically been associated with suboxic or anoxic
ground-water systems where reduction of Tc(VII) toTc(IV) is
favored. There is evidence that Tc(IV) may adsorb to solid
organic matter for systems undersaturated with respect
to precipitation of hydrous Tc(IV) oxides or other reduced
precipitate phases (Keith-Roach et al., 2003; Maes et al.,
2003). Due to the difficulty of differentiating adsorption of
Tc(IV) species versus precipitation or coprecipitation of
Tc(IV)-bearing solids, it is generally understood that appar-
ent retardation within an aquifer is due to the latter process.
(a)
o
111
-.2
-.4
-.6
TcO,
25
(b)
.2
111
-.2
-.4
10
12
14
pH
25°C
10
12
14
PH
Figure 5.1 Phase stability diagrams for technetium at
25°C. a) Eh-pH diagram forthe system Tc-H2O,
Tc= 10-75(53,903 pCi/L), showing metastable
field for TcO2-2H2O. b) Eh-pH diagram for
the system Tc-H2O, Tc = m10 to 1&8 (170-
17,045 pCi/L), all phases unsuppressed(LLNL
database).
46
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.6
(a)
.4
.2
Cfl
LU
-.2
-.4
-.6
TcO;
25° C
10
12
14
PH
.6
(b)
.4
.2
HI
-.2
-.4
-.6
0 2 4 6 8 10 12 14
PH
Figure 5.2 Phase stability diagrams for technetiumat25°C.
a) Eh-pH diagram for the system Tc-H2O-CO2,
Tc= 10-7S(53,903pCi/L), showing metastable
fieldforTcO2• 2H2O. Tc(IV)-carbonatecomplex
data from Hummel et al. (2002). b) Eh-pH
diagram for the system Tc-H2O-S; Tc = 10~75
and S=m3.
Site Characterization
Overview
Attenuation of "Tc might be achieved through precipita-
tion or coprecipitation reactions under reducing conditions.
Precipitation of sparingly soluble precipitates of Tc(IV) may
form under reducing conditions, but this would likely only
occur for relatively high concentrations of "Tc within the
plume. For example, it is generally reported that ground
water in which a hydrous Tc(IV) oxide controls aqueous
technetium concentrations will have dissolved "Tc con-
centrations on the order of 10~8 mole/L (17,045 pCi/L; Hu
et al., 2008). Coprecipitation with ferrous iron minerals
under iron- and/or sulfate-reducing conditions may be a
more likely pathway for immobilization. A list of potential
attenuation processes is provided in Table 5.1.
Aqueous Measurements
Under oxidizing conditions, "Tc will be present as the
soluble TcO4" anion, which is only slightly adsorbed to the
solid phase. As a beta particle emitter (Epirax= 294 keV), "Tc
can be quantified using scintillation counting techniques
(USEPA, 2004b; USEPA, 2006a). Interference from other
beta-emitting radionuclides, e.g., 90Sr (Epmax = 546 keV),
can be overcome through use of anion exchange resins
to selectively retain the pertechnetate anion. Beals et al.
(2000; 2001) and Fiskum et al. (2000) have demonstrated
the use of solid phase extraction (SPE) resins in permeable
disk configurations for rapid capture and analysis of "Tc in
water samples. However, Beals et al. (2001) have noted the
importance of re-measuring SPE disks at additional time
intervals to confirm that initial measured counts are not due
to interference for other short-lived beta-emitters that may
also be retained on the resin. In addition, initial tests for
extraction efficiency may need to be conducted to identify
whether major anion concentrations interfere with pertech-
netate uptake (e.g., elevated nitrate concentrations). There
have also been recent efforts to develop field-deployable
radiometric techniques to simplify and, in some cases,
automate measurements of "Tc in ground water (Egorov
et al., 2006; Hughes and DeVol, 2006; O'Hara et al., 2009).
A common configuration for these radiometric systems is
the physical mixing of an anion exchange resin and a solid
scintillator to allow detection of the emitted radiation using
a photomultiplier tube. A potential interfering radionuclide
is 129I, which is a beta emitter (Epmax = 294 keV) that may be
present in anionic form. However, O'Hara et al. (2009) has
demonstrated that the lower sensitivity for 129I detection will
limit its interference potential in plumes where it is present
in relatively lower concentrations than 99Tc.
Mass spectrometric techniques can also be used to mea-
sure "Tc in ground water (e.g., Hou and Roos, 2008).
Inductively coupled plasma-mass spectrometry (ICP-MS)
methods have been developed. The main interference in
determination of "Tc is mass overlap by atomic ions (99Ru)
or molecular ions (98MoH+) of mass 99, or peak tailing from
ions with mass in the range of 98-100 atomic mass units.
Several published studies illustrate use of ICP-MS for sen-
sitive and accurate detection of "Tc in complex aqueous
47
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Table 5.1 Natural attenuation and mobilization pathways for technetium.
Attenuation Processes
Mobilization Processes
Characterization Approach
Reduction of Tc(VII) and
precipitation of Tc(IV)
oxide/sulfide minerals
Oxidation of Tc(IV) to
Tc(VI) with formation of
pertechnetate anion
Evaluation of Tc concentration in ground water
and in solid matrix. Evaluation of Tc solid-phase
partitioning using sequential extraction methodologies
coupled to methods to determine Tc oxidation state.
Characterization of aqueous redox and chemical
conditions in ground water with speciation model
evaluation of potential Tc(IV) stability.
Reduction of Tc(VII) to
Tc(IV) and coprecipitation
with Fe(ll)-bearing
minerals
Oxidation of ferrous iron,
sulfide, and/or Tc(IV)
Evaluation of Tc concentration in ground water and in
solid matrix. Evaluation of Tc solid-phase partitioning
using sequential extraction methodologies; examine
correlation to extractable Fe, Ca, Mg and S. Batch and
column testing to determine Tc uptake behavior and
capacity of site-specific aquifer materials under variable
geochemical conditions.
matrices (e.g., Richteret al., 1997; Eroglu etal., 1998; Kim
et al., 2002). The potential for carryover of "Tc on system
components between samples needs to be evaluated, but
methods are available to eliminate this factor (e.g., Richter et
al., 1997). Due to similarities in chemical behavior, rhenium
(as perrhenate anion) can be used to monitor chemical
recovery of pertechnetate during sample pre-concentration
prior to detection (Mas et al., 2004).
Solid Phase Measurements
Solid phase measurements that may provide information
useful to assessing processes controlling "Tc attenuation
within the aquifer and the capacity along relevant transport
pathways include the determination of "Tc partitioning to
aquifer solids and identification of aquifer solids mineral-
ogy that may participate in reduction or precipitation reac-
tions. Evaluation of the mass distribution of "Tc between
co-located ground water and aquifer solids throughout the
plume provides an assessment of the extent that retarda-
tion reactions limit strontium migration. Bulk solid-phase
partitioning can be conducted using total digestion or
acid-extractable techniques with appropriate measures
to assess extraction efficiency and potential interference
from other beta-emitters that may interfere with radiometric
techniques (USEPA, 2004b; USERA, 2006a). Because the
overall quantities of solid-phase "Tc are likely to be low,
techniques for determination of total "Tc or the chemical
fractionation of "Tc in aquifer solids is needed. Fusion
methods that convert the solid matrix to a form easily dis-
solved in acid are recommended for total concentration
measurements (Dixon et al., 1997; Zhao et al., 2008). The
fusion methods incorporate reagents that minimize volatile
losses of technetium.
The relative partitioning of "Tc with mineral components in
the aquifer solids may be inferred from extractions designed
to selectively dissolve reactive phases (e.g., carbonates,
reducible oxides, or oxidizable sulfides). This approach has
been used to assist interpretation of "Tc speciation in sedi-
ments (Keith-Roach et al., 2003). However, interpretation
of sequential extraction data needs to be constrained by
knowledge of the mineralogical composition of the aquifer
solids, the types of major elements also extracted in any
particular step, and the geochemical conditions under which
the solids existed within the subsurface. Additional infor-
mation on analysis approaches and analytical techniques
applied to solid phase characterization is provided in USEPA
(2007; Section NIB). There are no methods that have been
developed to identify the oxidation state of extracted "Tc at
the concentrations likely to be encountered in aquifer solids.
It should also be noted that dissolution of aquifer solids
will likely release elements that interfere with radiometric
or mass-spectrometric detection of "Tc at concentrations
much higher than typically encountered in ground water.
The performance of analytical procedures used to quantify
released "Tc needs to be carefully evaluated in the context
of the complex liquid matrices that will be encountered. The
oxidation state and chemical binding environment of "Tc
may also be assessed using X-ray absorption spectroscopy
(see references in Solubility section). This technique can
be used to identify the chemical speciation of technetium
in aquifer solids collected and preserved in a manner to
maintain in-situ characteristics.
Long-term Stability and Capacity
Two factors that will likely dictate the stability of attenu-
ated "Tc and the capacity for the aquifer to sustain "Tc
removal from ground water include the persistence of
48
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reducing conditions and the mass flux of "Tc transported
along relevant flow paths. Identification of contaminant
source areas, including waste form characteristics, may
be necessary in order to understand contaminant loading
relative to the source and flux of water moving through the
plume. For example, it may be useful to examine isotopic
signatures for other fission-product isotopes that provide
unique markers for different contaminant sources (e.g.,
fission-produced stable ruthenium isotopes; Brown et
al., 2006). Characterization of the total mass and rate of
release of "Tc from uncontrolled source areas need to be
understood relative to the apparent rates of pertechnetate
reduction and sequestration in the plume. The apparent
rate of attenuation at the leading edge of the plume can
be assessed, by reference to a conservatively transported
constituent (Hu et al., 2008), through time-series sampling
at individual wells or by sampling from a well transect
installed coincident with the direction of water flow. A situ-
ation where the flux of "Tc entering the plume exceeds the
rate of attenuation at the leading edge of the plume will likely
result in insufficient capacity to arrest plume expansion. The
factors governing reducing conditions in the plume may also
govern attenuation capacity, particularly in situations where
the flux of degradable organic compounds or other electron
donors through the plume may not be sustained. However,
examination of the stability of reduced forms of technetium
sequestered to subsurface solids suggests that solid phase
Tc(IV) will remain stabile, even if more oxidizing conditions
develop (e.g., McBeth et al., 2007; Begg et al., 2008).
Tiered Analysis
Determination of the viability of "Tc remediation in ground
water via monitored natural attenuation will depend upon
proper assessment of contaminant loading to the aquifer
and prevailing geochemistry and mineralogy within the
contaminant plume and the down gradient zone prior to
the point(s) of compliance. Due to the long half-life for
"Tc, radioactive decay will not provide a viable mechanism
for plume attenuation. Therefore, the goal of site assess-
ment will be to demonstrate the process(es) controlling
"Tc sequestration onto aquifer solids and the long-term
stability of solid phase "Tc as a function of existing and
anticipated ground-water chemistry. The following tiered
analysis structure for site characterization provides a techni-
cally defensible approach to evaluate candidate sites and
define the potential limitations of MNA as part of a remedy
for ground-water cleanup.
Tier I. Site characterization under Tier I will involve demon-
stration that the ground-water plume is static or shrinking,
has not reached compliance boundaries, and does not
impact existing water supplies. Once this is established
through ground-water characterization, evidence is col-
lected to demonstrate "Tc partitioning to aquifer solids
within the plume. If natural attenuation processes are active
throughout the plume, then there should be an observed
increase in solid phase concentrations within regions
of the plume with higher aqueous concentrations, e.g.,
near the source term. This field partitioning data may be
supplemented by geochemical modeling that incorporates
measured water chemistry (e.g., pH, Eh, and major ion
chemistry) throughout the plume to assess the potential
for solubility control by a "Tc precipitate such as a hydrous
oxide phase. Identification of active sequestration to pre-
vent "Tc migration in ground-water provides justification
for proceeding to Tier II characterization efforts.
Tier II. Under Tier II, the apparent rate and mechanism(s)
of attenuation are determined. Estimates of a site attenu-
ation rate(s) can be assessed via a well transect along
the ground-water flow path. In addition, time-series data
may be collected at one or more monitoring points within
the plume (USEPA, 2007; Section IIIA.5). This informa-
tion will allow assessment of the relative timescales for
contaminant immobilization and fluid transport and deter-
mination of whether remediation objectives can be met
within the required regulatory time frame. In addition, the
mechanism(s) for attenuation are to be identified under
this stage of site characterization. This effort will require
determination of the chemical speciation of solid phase "Tc
and may be approached according to the following scheme:
1) Determination of solution and solid phase "Tc
concentrations, along with the relative concentration
of major ions/components in aquifer solids where
attenuation is occurring;
2) Calculation of saturation state of ground water
relative to measured aqueous chemistry;
3) Determination of aquifer mineralogy (Amonette,
2002) to determine the relative abundance of
components that might support pertechnetate
reduction (e.g., Fe(ll) associated with aquifer solids)
and/or coprecipitation; and
4) Determination of "Tc-sediment associations via
chemical extractions designed to target specific
components within the aquifer solids.
This compilation of information will facilitate identification of
the reaction(s) leading to "Tc immobilization. It is recom-
mended that identification of redox-sensitive components
in aqueous and solid matrices be conducted using samples
collected in a manner that preserves their in-situ specia-
tion (USEPA, 2006b). The demonstration of concurrence
between conceptual and mathematical models describing
"Tc transport will entail development of site-specific param-
eterization of the chemical processes controlling "Tc solid
phase partitioning.
Tier III. Once the partitioning mechanism(s) have been
identified for the site, the subsequent characterization effort
under Tier III will involve determination of the stability of
immobilized "Tc and the capacity of the aquifer to sustain
continued uptake. It is recommended that the stability of
immobilized "Tc be tested based on the anticipated evo-
lution of ground-water chemistry concurrent with decay of
the plume. For example, changes in ground-water pH can
exert a significant influence on the solubility of a hydrous
Tc(IV) oxide. Therefore, it is recommended that sediment
leach tests be conducted to characterize the magnitude
of "Tc mobilization as a function of pH for a ground-water
49
-------
chemistry representative of site conditions. It is recom-
mended that the capacity for "Tc uptake onto aquifer solids
be determined relative to the specific mechanism(s) identi-
fied in Tier II. For example, if site characterization under
Tier II indicated that microbial degradation of naturally
occurring solid organic matter (SOM) resulted in reduction
of Tc(VII) to insoluble Tc(IV), then the mass distribution
of SOM within the aquifer needs to be determined. This
site-specific capacity would then be compared to "Tc mass
loading within the plume in order to assess the longevity of
the natural attenuation process. If site-specific tests dem-
onstrate the stability of immobilized "Tc and that there is
sufficient capacity within the aquifer to sustain "Tc attenu-
ation, then the site characterization effort can progress to
Tier IV. For cases where contaminant stability is sufficient
but aquifer capacity is insufficient for capture of the entire
plume, then a determination of the benefits of contaminant
source reduction is required.
Tier IV. Finally, under Tier IV a monitoring plan is estab-
lished along with contingency plans in the event of MNA
failure. It is recommended that the monitoring plan be
designed to establish both continued plume stability and
to identify changes in ground-water chemistry that may
lead to re-mobilization of attenuated "Tc. The specific
chemical parameters to be monitored will include those
identified under Tier III that may halt "Tc partitioning and/or
result in dissolution of either discrete "Tc precipitates or
aquifer minerals that sequester "Tc from ground water.
For example, solution phase parameters that could alter
either "Tc precipitation include inorganic carbon (alkalinity),
pH, and dissolved oxygen. In contrast, increases in the
concentration of sulfate may indicate the dissolution of an
important sorbent phase within the aquifer (e.g., oxidative
dissolution of iron sulfide). Changes in these parameters
may occur prior to observed changes in solution "Tc and,
thus, serve as monitoring triggers for potential MNA failure.
In this instance, a contingency plan can be implemented
that incorporates alternative strategies to arrest possible
plume expansion beyond compliance boundaries. Possible
strategies to prevent plume expansion include pump and
treat operations, installation of reactive barriers to enhance
uptake capacity perpendicular to the direction of plume
advance, or enhancement of attenuation processes within
the aquifer through the injection of soluble reactive com-
ponents that induce more reducing conditions.
References
Abdelouas, A., B. Grambow, M. Fattahi, Y. Andres, and E.
Leclerc-Cessac. Microbial reduction of "Tc in organic
matter-rich soils. Science of the Total Environment
336:255-268 (2005).
Amonette, J.E. Methods for determination of mineralogy
and environmental availability, p. 153-197, In J.B.
Dixon and D.G. Schulze (eds.) Soil Mineralogy with
Environmental Applications, SSSA Book Series, no. 7.
Soil Science Society of America, Madison, Wl (2002).
Beals, D.M., B.S. Crandall, and RD. Fledderman. In-situ
sample preparation for radiochemical analyses of
surface water. Journal of Radioanalytical and Nuclear
Chemistry 243:495-506 (2000).
Beals, D.M., K.J. Hofstetter, V.G.Johnson, G.W. Patton, and
D.C. Seely. Development of field portable sampling
and analysis systems. Journal of Radioanalytical and
Nuclear Chemistry 248: 315-319 (2001).
Beaver, J.E. and H.B. Hupf. Production of "mTc on a
medical cyclotron: A feasibility study. Journal of Nuclear
Medicine 12:739-741 (1971).
Begg, J.D.C., IT. Burke, J.M. Charnock, and K. Morris.
Technetium reduction and reoxidation behaviour in
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52
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Uranium
James E. Amonette, Richard T. Wilkin, Robert G. Ford
Occurrence and Distribution
Uranium is the heaviest naturally occurring element and all
isotopes are radioactive. Although it occurs as an essential
component in nearly 200 different minerals (Smith, 1984;
Burns, 1999), the vast majority of naturally occurring ura-
nium deposits are comprised of a few common minerals
including oxides (uraninite and pitchblende), silicates (cof-
finite, soddyite, uranophane, and uranothorite), phosphates
(autunite), and vanadates (carnotite). Uranium can be
used to fuel nuclear power reactors either in its natural
isotopic enrichment of 0.7% 235U or at slight enrichments
of between 3 - 5% 235U. At much higher 235U enrichments,
uranium is also used in nuclear weapons. Civilian non-
nuclear applications of (depleted) uranium include sailboat
keels and counterweights in commercial aircraft. Depleted
uranium is used extensively in military applications, notably
in alloys for armor and armor-penetrating projectiles (Bleise
et al., 2003). As a result of its widespread use, uranium
is the most common radiological contaminant in soils and
sediments (Riley et al., 1992) and is generally associated
with uranium mining/refining sites, processing facilities for
the commercial/civilian nuclear fuel cycle, production of
nuclear materials for weapons, and artillery firing ranges
and battlefields.
Typical groundwater concentrations of dissolved uranium
are on the order of a few ug U/L, but range as high as
2000 ug U/L near natural uranium deposits in arid regions.
In streams, levels of uranium are typically about 1 ug U/L,
but values as high as 50 ug U/L are seen in granitic water-
sheds. The level of uranium in seawater is 3 ug U/L. In
addition to the dissolved forms, colloidal forms of uranium
are often present in water, and can be determined by
decreases in concentration as a result of filtration (the filter
pore size is often specified).
Some of the notable uranium contamination sites in the
United States include Department of Energy (DOE) pro-
cessing facilities located at Fernald, OH; Paducah, KY;
Oak Ridge, TN; Savannah River Site, SC; Rocky Flats,
CO; and Hanford, WA, as well as abandoned mines and
former mill sites in the Colorado Plateau region of the
southwest. In addition, the development of in-situ leach
mining, which accounts for nearly all the uranium currently
mined in the United States, has led to new sites of sub-
surface contamination in Texas, Wyoming, Nebraska, and
New Mexico. Military firing ranges and research facilities
contaminated with depleted uranium include the Aberdeen
Proving Ground, MD; Jefferson Proving Ground, IN;Yakima
Firing Center, WA; and Los Alamos and Sandia National
Laboratories, NM. A summary of sites under the regulatory
authority of either the USEPA or the Nuclear Regulatory
Commission (NRC) was provided in USEPA (1993).
Geochemistry and Attenuation Processes
Radioactive Decay
Uranium transport in groundwater depends to a large extent
on the oxidation state and radioactive decay phenomena
for the predominant radioisotopes found in natural sys-
tems (USEPA, 1999). All of the isotopes of uranium are
radioactive. Of the three naturally occurring isotopes, 238U
is the most abundant (99.275%; 4.468x109 year half-life),
followed by 235U (0.720%; 7.038x108 year half-life) and 234U
(0.005%; 2.445x105 year half-life). The decay series for 238U
and 235U are shown in Figure 6.1. Because 234U is a decay
product of 238U, it generally occupies sites in minerals that
have already been damaged by the release of an alpha
particle during the 238U decay process. This decay process,
referred to as 'alpha recoil', may lead to ejection of the
238U daughter (234Th) directly into groundwater or indirectly
increase the potential for 234U leaching from aquifer solids
damaged by the recoil process (Ivanovich, 1994; Tricca
et al., 2001; Maher et al., 2006). The 234U:238U activity ratio
in groundwater would increase proportionately, where 234U
is released during dissolution or from the rapid decay of
234Th (and 234Pa) following ejection from the solid surface.
The form of uranium waste materials will influence the
relative distribution of uranium radioisotopes derived from
sources of contamination. Preparation of uranium for use
in reactors involves enrichment of the more fissile 235U
radioisotope to several percent, and for nuclear weapons
enrichment to 90% or better is required. The residual
material, or tails, derived from the enrichment process
is referred to as "depleted uranium", which has a higher
proportion of the 238U radioisotope (Meinrath et al., 2003).
Elevated levels of 236U are produced by neutron activation
of 235U (Marsden et al., 2001; Boulyga and Becker, 2002);
accordingly, this radioisotope may provide a "fingerprint" for
the presence in the environment of uranium that has been
irradiated in a nuclear reactor (Christensen et al., 2007) due
to its relatively long half-life (2.3415 x 107 years). Thus,
excluding potential alpha recoil effects, differences in the
relative abundance of uranium isotopes in groundwater will
reflect source term characteristics. Possible progenitors
that could introduce uranium isotopes via decay in-growth
are shown in Table 6.1. Radioisotopes of neptunium,
protactinium, and plutonium decay to produce uranium
radioisotopes of importance. The decay half-lives for these
progenitors range from minutes to many thousands of years.
Knowledge of the presence of these radionuclides within
the plume and sampled groundwater may be important for
proper identification of the source of uranium radioisotopes
as well as in-growth corrections that may be required to
properly account for the mass/activity of the various uranium
radioisotopes at the time of sample collection.
53
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Table 6.1 Illustration of potential decay paths from different progenitor sources leading to production of uranium ra-
dioisotopes. Determination of possible decay paths to the target radionuclide was based on examination
of the Chart of Nuclides (http://www.nndc.bnl.gov/chartA maintained by the Brookhaven National Labora-
tory, National Nuclear Data Center (NNDC) relative to possible decay paths based on decay modes identi-
fied in the Appendix (EC = electron capture, p" = electron emission, a = alpha decay). Decay half-life data
were obtained using the WinChain program that provides electronic access to the ICRP38 Nuclear Decay
Data Files (ICRP, 1983; Eckerman et al., 1994; m = minutes, h = hours, d = days, y = years). WinChain is
a public domain software application available for download from Oak Ridge National Laboratory
(http://ordose. ornl. gov/downloads. html).
Contaminant Radionuclide
238(J
235(J
234(J
236U
Decay Progenitor
242pu
235Np
239Pu
234Np
234pa
238pu
238Np^238pu
236Np 1
240pu
Decay Mode
a
EC
a
EC
P-
a
P~-»a
EC
a
Progenitor Decay Half-life
3.763 x 105y
396.1 d
24065 y
4.4 d
6.70 h
87.74 y
2.1 17 d-» 84.74 y
1 53000 y
6537 y
Decay data not available in WinChain; recommended values obtained from NuDat 2.4 database maintained at the NNDC
(http://www. nndc. bnl. gov/nudat2/index.jsp).
54
-------
7.04110 *y
"'Pa
3.28x10*y
227 Ac
21. By
a
18.7 d
(0.988)
a
223 Ra
11.4 d
^Fr I
22.00m •
(0.014} !
a
2ieRn
396>
a
215Po
0.00178s
a
2,1 Rb
36.1 m
'Bi
p-
a
207JI
477m
(0.9672}
211 Po
0.521
_ (0,0028)
a
207 Pb
slot*
Figure 6.1 Decay series for 238U and 235U based on data from ICRP (1983). Decay modes include those leading to
ejection of an alpha particle (a) or a beta particle (p~). The half-life is shown directly below the isotope that
is subject to decay;y = years, d = days, h = hours, m = minutes, s = seconds, \js = microseconds. Numbers
shown in parentheses below indicate the fractional abundance of the daughter isotope produced during
branched decay of the parent isotope (e.g., decay of218Po follows two routes resulting in 98.98% production
of214Pb and 0.02% production of218At).
55
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Aqueous Speciation
The environmental chemistry of uranium is largely dictated
by its formal oxidation state (e.g., Fanghanel and Neck,
2002). Under ambient oxidizing conditions, the predomi-
nant uranium oxidation state is U(VI). Where oxygen is
limited, U(IV) may dominate. The metallic form, U(0), does
not occur naturally, and is readily oxidized to U(IV), and
eventually U(VI), upon exposure to oxidizing conditions. A
considerable body of literature discusses the mechanisms
of "corrosion" of U(0) and U(IV) to U(VI) (e.g., Finch and
Ewing, 1992; Finch and Murakami, 1999). Other oxidation
states of uranium, e.g., U(V)and U(lll), are rare and gener-
ally unstable to U(IV) and U(VI) under ambient conditions.
In general, the solubility, and hence mobility, of uranium
is greatest when it is in the U(VI) state. Complexation of
U(VI) by inorganic anions such as carbonate, fluoride and
phosphate may enhance the solubility and mobility of this
species. When reducing conditions are present, U(IV) is
generally immobile and found either as the insoluble oxide
(uraninite) or the silicate (coffinite). A compilation of ther-
mochemical data for uranium aqueous species and solid
phases has been recently published by the Nuclear Energy
Agency (Guillaumont et al., 2003). It is recommended that
this and other sources (e.g., Hummel etal., 2002; Gorman-
Lewis et al., 2008) be consulted prior to implementation of
geochemical models to describe uranium fate and transport.
Under oxidizing conditions and environmental pH, U(VI)
species dominate aqueous uranium concentrations. These
highly soluble species are generally either hydroxy or car-
bonato complexes of the uranyl (UO22+) cation (Figure 6.2),
although elevated concentrations of potential inorganic or
organic ligands near contaminant source zones may exert
greater influence on U(VI) speciation (e.g., phosphate;
Bonhoure et al., 2007). As shown in Figure 6.2, calcium
(or other alkaline earth metals such as magnesium) and
inorganic carbon in ground water tends to dominate the
aqueous speciation of U(VI) under typical pH conditions
(Dong et al., 2005; Dong and Brooks, 2006; Kelly et al.,
2007; Dong and Brooks, 2008). The presence of these
species at moderate ground-water concentrations has been
verified, along with the reliability of the stability constants for
calcium-carbonato complexes of U(VI) (Prat et al., 2009).
As noted below, these speciation characteristics also influ-
ence the degree to which U(VI) will adsorb to aquifer solids.
Under reducing conditions, U(IV) species, primarily the
uranyl cation and its complexes, predominate, but, due to
the very low solubility of U(IV) minerals, reach maximum
concentrations on the order of 10 nM (2.4 ug U/L). For all
practical purposes, therefore, only U(VI) aqueous species
are at sufficient concentrations to be of environmental
concern. Complexation with dissolved organic carbon may
influence the aqueous chemistry of U(VI) in some ground-
water systems, although field evidence suggests that this
would be significant only in more acidic (pH<6) systems
(e.g., Ranville et al., 2007).
100 -
s- 80 -|
§
•2 60 -
40 -
O 20 -
01
0 -
(a) pH 7
•O- OxyHydroxyl
^^^ Sulfato
•^- Carbonate
Calcium carbonato
I ' I " '
0.000 0.002 0.004 0.006 0.008 0.010
Alkalinity (mol/L CaCO3)
(b) 100 mg/L CaCO
3(aq)
- 100
-60
-40
-20
- 0
CD
7
PH
Figure 6.2 Distribution of aqueous U(VI) species among oxyhydroxyl-, sulfate-, carbonate-, and calcium carbonate-
complexes as a function of ground-water chemistry, (a) Dependence on alkalinity expressed as mol/L
CaCO3(aq) ;pH = 7, 0.005 mol/L NaCI, 0.001 mol/L K2SO4, 0.001 mol/L MgNO3 0.42 \jmol/L U (100 ug U/L).
(b) Dependence on pH; 0.001 mol/L CaCO3( AWOmg/L CaCO3alkalinity), 0.005 mol/L NaCI, 0.001 mol/L
K2SO4, 0.001 mol/L MgNO3, 0.42 \\mol/L U (100 ug U/L). Model predictions using Visual MINTEQ
Version 2.53 (Based on MINTEQA2 described in Allison et al. (1991); available at http://www.Iwr.kth.se/
English/OurSoftware/vminteq/). Precipitation of Ca/Mg carbonates suppressed; slight oversaturation for
pH>8 and/or> 0.005 M CaCO3(aq}.
56
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Solubility
Under ambient conditions, the thermodynamically stable
uranium solid phases will be either U(VI) or U(IV) com-
pounds. Thus, depleted U(0) will oxidize to U(IV) and, if
sufficiently oxidizing conditions pertain, eventually to U(VI).
The most stable U(VI) compounds are the phosphates
and vanadates, but their formation is often limited by the
relatively low concentrations of these two anions, and thus
more soluble U(VI) oxides such as schoepite, which is bright
yellow in color, are often seen, if any U(VI) solid phases are
present. A significant fraction of the solid-phase U(VI) will
be adsorbed to iron (hydr)oxide surfaces, the edges of clay
minerals, and to organic matter, rather than precipitated as
discrete U phases (discussed below). Maximum solubility of
uranium is seen in oxidizing, phosphate-free, carbonate-rich
solutions, and these are the principal reagents used for in-
situ leach mining of uranium in the U.S. There are examples
of uranium substitution into calcites from unimpacted
(Sturchio et al., 1998; Kelly et al., 2003; Kelly et al., 2006) or
contaminated environments (Wang et al., 2005a; Catalano
et al., 2006), but this generally appears to occur only in
locations where active precipitation of calcite or aragonite
is occurring. The stability fields of U(VI) as a function of pH
and various element combinations are shown in Figures 6.3
and 6.4. The pH-dependent solubility of U(VI) is shown in
Figure 6.3 for systems containing carbonate and silica. In
Figure 6.4, the role of phosphate in controlling U(VI) con-
centrations is illustrated. Precipitates of U(VI) have been
observed to form near source areas where elevated uranium
concentrations in ground water were likely to have existed.
An example where the U(VI) silicate, sodium boltwoodite,
formed near a uranium source area has been documented
for the Hanford Site in Richland, WA (Catalano et al., 2004;
McKinley et al., 2007). Examples of the types of uranium
phosphate precipitates that have been identified include
uranyl phosphate and uranium metaphosphate (Buck et
al., 1996; Morris et al., 1996), barium meta-autinite (e.g.,
Jerden et al., 2003; Jerden and Sinha, 2003; Jerden and
Sinha, 2006), and metatorbernite (Catalano et al., 2006;
Aral et al., 2007).
Under reducing conditions, the stable U(IV) solid phases
are uraninite and, if high dissolved silica pertains, coffinite.
Organic complexes of U(IV) associated with humic material
may also retain U(IV) in the solid phase. The solubility of
the U(IV) phases is extremely low, and thus the presence
of reducing conditions effectively halts the movement of
uranium in soils and sediments, provided that colloidal-
sized phases are not formed and transported. The most
common uranium ore-forming process involves reductive
precipitation of U(IV) phases as a result of microbiologi-
cal activity to form a roll-front deposit (Langmuir, 1997).
The stability fields for U(VI) and U(IV) as a function of
pH and Eh for various water compositions are shown in
Figure 6.5. The geochemical modeling results presented
in Figures 6.3, 6.4, and 6.5 suggest that a wide variety of
uranium-bearing precipitates are possible, especially in
complex ground-water systems that invariably contain silica,
carbonate/bicarbonate, calcium/magnesium, sodium, and
sometimes phosphate. Furthermore, it may be difficult to
predict associations of uranium in the solid phase based
upon analysis of aqueous chemical data and solubility
predictions from thermochemical data. In the absence of
confirmatory solid phase characterization data, equilibrium
model projections only indicate the possible formation of
specific uranium-bearing precipitates.
Figure 6.3 (a) Solubility of U(VI) as a function of pH at
various levels of PCO2 and at aH4SiO4=1Q-40.
Bold lines show the pH-dependent solubility
ofsoddyite[(UO2)2SiO4-2H2O\ atPCO2=W30
bar. Note that the stability field of soddyite
decreases with increasing PCO2. Dashed
lines show the metastable extensions of U( VI)
aqueous species and the pH-dependent solu-
bility of schoepite[ f,-UO3-2H2O] atPCO2=ia3
bar. (b) Addition of Ca (ia4 to ia2) and
A/a (10-3). Thermodynamic data for urano-
phane(Ca(H30)2(U02)2(Si04)2-3H20)arefrom
Langmuir (1997; p. 552). Note that solubility
studies by Perez et al. (2000) suggest that
uranophane maybe less stable than depicted
here.
57
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a)
9 1C
Figure 6.4 Solubility of U(VI) as a function of pH and in
the presence of phosphate (1Q-35), sodium
(10-3), and calcium (10-3). Na-Autunite
(Na2 (UO2 )2 (PO4 )2) is the least soluble solid
across the pH range, followed by Autunite
(Ca(U02 )2 (P04 )2), (U02 )3 (P04 )2 -4H2 O(c),
Schoepite, and UO2 HPO4 -4H2 O(c). Dashed
lines show the metastable extensions of aque-
ous U(VI) species. Thermodynamic data for
U-P complexes and solids are from Guillaumont
etal. (2003).
PH
Figure 6.5 Eh-pH diagrams for U at 25°C. (a) System
U-O-H-Si-C, with aH4 SiO4 = 0~4° and
PCO2 = 10-25 bar. (b) System U-O-H-C
with EL/ = 10-s and PCO2 = 10~2S bar.
(c) System U-O-H-C-Na-Ca with EL/ = 1 a5,
aH4 SiO4 = 10-40, and PCO2 = 10~25 bar.
Soddyite = (UO2 )2 SiO4 -2H2 O; urani-
nite = UO2; schoepite = f>-UO3 -2H2 O; urano-
phane = Ca(H3 O)2 (UO2 )2 (SiO4 )2 -3H2 O.
58
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Adsorption
Adsorption of uranium typically involves inner-sphere com-
plexation of uranyl (i.e., those containing UO22+) species
by oxygen ligands at the surfaces of iron oxyhydroxides,
phosphates, and layered silicates. Uranyl species exhibit a
high affinity for iron oxyhydroxide surfaces and for both basal
and edge sites on layered aluminosilicates such as smectite
and vermiculite. Adsorption of U(VI) to the aluminosilicate
mineral, muscovite, has been observed in aquifer sediments
at the Hanford Site in Richland, WA (Catalano et al., 2006;
McKinley et al., 2007). Complexation of U(VI) by organic
ligands in solid humic materials (primarily carboxylic-acid
and phenolic groups) may also serve to remove uranium
in shallow ground-water systems (Sowder et al., 2003).
A compilation of published Kd values for U(VI) sorption onto
soils/sediments is documented in USEPA (1999). However,
as recognized by the authors of this compilation, there are
significant limitations to the application of published Kds for
site-specific applications where either the ground-water
chemistry or the aquifer matrix differs significantly from the
conditions under which a Kd was determined (Ochs et al.,
2006). Davis and others (Davis and Curtis, 2003; Davis
et al., 2004; Kohler et al., 2004) document an alternative
approach whereby a site-specific Kd value is modeled
through the use of a nonelectrostatic surface complexation
model (NEM) developed as a function of site geochemis-
try for aquifer sediments. This approach incorporates the
important influence of uranium solution speciation while
avoiding the need to model the influence of individual
mineral components (and their respective surface charging
behavior). While this approach still requires site-specific
data, it provides a means for projecting the influence of
changes in ground-water chemistry on uranium sorption.
The chemistry of ground water may be influenced by reac-
tion with aquifer solids and/or external recharge/infiltration
from atmospheric precipitation or surface water. As previ-
ously noted, alkalinity influences the aqueous speciation of
U(VI), and it also influences the degree of sorption of U(VI)
onto iron oxyhydroxides (e.g., Waite et al., 1994) and aquifer
solids in which these minerals control uranium partition-
ing (e.g., Sato et al., 1997; Urn et al., 2007). It has been
demonstrated that changes in ground-water chemistry influ-
ence the transport of U(VI) through an aquifer (Dong et al.,
2005; Yabusaki et al., 2008). Alternatively, transition from
oxidizing to reducing conditions along the transport pathway
may be accompanied by a shift from adsorption of U(VI)
species to precipitation of U(IV)-bearing solids (Davis et
al., 2006). Thus, it is recommended that reactive-transport
models used to project subsurface uranium mobility directly
incorporate the influence of major ion chemistry and redox
conditions on the chemical speciation of uranium.
There is field evidence that adsorption of uranium to
mineral surfaces within an aquifer may be an intermedi-
ate step to the formation of uranium-bearing precipitates.
Murakami et al. (1997; 2005) have observed the association
of nanoparticulate U(VI)-phosphate precipitates with iron
oxyhydroxides in the weathering zone downgradient from
a uranium ore deposit. The U(VI) mineral was identified as
metatorbernite, which was present in ground water that was
undersaturated with respect to precipitation of this mineral.
Characterization of the textural associations between the
nanocrystalline metatorbernite and iron oxyhydroxides
present as fissure fillings, clay coatings, and nodules,
along with compositional relationships between copper,
phosphorous, and uranium (Sato et al., 1997) indicated that
the formation of uranium precipitates was a secondary step
following initial adsorption of these constituents onto iron
oxyhydroxide mineral surfaces (Murakami et al., 2005). As
summarized by Payne and Airey (2006), the observations
in this subsurface system provide a point of reference for
designing site characterization strategies and developing
both conceptual and analytical models for interpreting and
projecting uranium mobility in ground water.
Site Characterization
Overview
Uranium mobility in ground water is governed by the total
concentration of uranium, the distribution of uranium spe-
cies in water, and the nature of uranium partitioning in the
solid phase. The development of conceptual site models
for predicting the long-term fate of uranium at a contami-
nated site will require information on the concentration and
chemical speciation of uranium in the aqueous phase and
the solid phase. Table 6.2 illustrates possible attenuation
and mobilization pathways for uranium in ground water.
Details of the types of analytical measurements that may
be conducted on sampled ground water and aquifer sedi-
ments to assist in identifying the attenuation mechanism(s)
are discussed in the following paragraphs.
Aqueous Measurements
Overviews of radiometric techniques for determining the
activity of uranium radioisotopes are provided in Tosheva
et al. (2004) and USEPA (2006a). The sensitivity of these
methods are generally good for uranium radioisotopes, but
typically isolation of the analyte from the sample matrix is
required prior to analysis. Becker (2003) and Lariviere et
al. (2006) provide recent reviews of mass spectrometry
applications for the determination of radionuclide concen-
trations in environmental samples. The sensitivity and
mass-selectivity of these approaches, along with the ability
to circumvent matrix and isobaric interferences, have sig-
nificantly increased the utility of these methods. This can
be further improved with the use of on-line techniques for
the separation and enrichment of the targeted radionuclide
from the sample matrix (e.g., Unsworth et al., 2001), similar
to approaches used for radiometric measurements. Recent
advances in automation of steps to isolate and concentrate
the analyte prior to introduction to plasma-based mass
spectrometers has increased the speed and utility of these
multi-element analytical techniques.
Metilda et al. (2007) described development of an elec-
trochemical sensor that facilitates selective extraction
and detection of the uranyl anion [U(VI)j from natural
water samples with a 5 |ag/L detection limit. The sensor
is constructed using an ion implanted polymer that can
59
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Table 6.2 Natural attenuation and mobilization pathways for uranium.
Attenuation Processes
Mobilization Processes
Characterization Approach
Precipitation of uranium
as U(IV) oxide or silicate
phases in reduced
systems
Increase in pH or alkalinity
and/or transition to oxidizing
conditions
Evaluation of U concentration in ground water
and in solid matrix. Evaluation of U solid-phase
partitioning using sequential extraction methodologies
coupled to methods to determine U oxidation state.
Characterization of aqueous redox and chemical
conditions in ground water with speciation model
evaluation of potential U(IV) stability.
Precipitation of uranium
as U(VI) silicate or
phosphate phases in oxic
systems
Increase in alkalinity or
dissolved organic carbon
capable of forming solution
complexes; decrease
in silicate/phosphate
concentration
Evaluation of U concentration in ground water and in
solid matrix. Evaluation of U solid-phase partitioning
using sequential extraction methodologies coupled
to methods to determine U oxidation state; examine
correlation to extractable phosphate and/or alkaline
earth metals. Characterization of aqueous redox and
chemical conditions in ground water with speciation
model evaluation of potential U(VI) stability.
Adsorption or
coprecipitation of U(VI)
with iron oxyhydroxides,
iron sulfides, and
carbonates or adsorption
onto clay mineral surfaces
Desorption due to high
pH, high competing anion
concentrations (e.g.,
carbonate), or high DOC
concentrations. Reductive
dissolution of iron hydroxides
or oxidative dissolution of
iron sulfides.
Evaluation of U concentration in ground water and in
solid matrix. Evaluation of U solid-phase partitioning
using sequential extraction methodologies; examine
correlation to extractable Fe, Ca, Mg and S. Batch
and column testing to determine U uptake behavior
and capacity of site-specific aquifer materials under
variable geochemical conditions.
be incorporated into optical detection configurations for
screening or quantification of uranyl concentrations in aque-
ous solutions (James et al., 2008; Prasada Rao and Kala,
2008). The equipment needed for use of these sensors
is suitable for use in field measurements. Recently, Liu et
al. (2007) has demonstrated synthesis of a UO22+-specific
DNAzyme that can be incorporated into a simple sensor
configuration for selective and sensitive quantification of
uranyl. The authors demonstrated the performance of
this analysis method for bicarbonate extracts of uranium-
contaminated soils. Alternatively, detection of the uranyl
dioxo-cation (UO22+) can also be achieved in the laboratory
using methods such as laser-induced kinetic phospho-
rimetry (Brina and Miller, 1992; Sowder et al., 1998; Elias
et al., 2003), time-resolved laser-induced fluorescence
spectroscopy (Wang et al., 2004;TRLIFS), or time-resolved
emission spectroscopy (Billard et al., 2003). The latter
two methods can be used to determine if multiple uranium
species make up the mobile form of uranium. Greene et
al. (2005) describe a method based on ultraviolet-visible
detection of U(VI) following complexation with the organic
chelating agent arsenazo III. Their adaptation of using
solid phase extraction with C18 resin following buffering of
the sample in pH 2 malonic acid resulted in elimination of
interferences from commonly occurring ions and a detec-
tion limit of 40 nanogram/L. Thus, in addition to traditional
methods that have been employed (e.g., USEPA, 2006a),
there are a range of available methods for quantification of
the uranyl ion in ground water at low concentrations, which
will facilitate determination of the chemical speciation of
mobile uranium when combined with analysis of the total
uranium concentration.
Determination of the distribution of uranium radioisotopes
may also be needed to determine the potential source
or sources of uranium contributing to the ground-water
plume. As outlined previously, contaminant sources of
uranium from the production or processing of uranium
for nuclear energy applications will lead to materials that
are enriched or depleted in 235U (or enriched in 236U for
reprocessed materials). Characterization of the ratios of
uranium radioisotopes (^\J:23B\J, sas^ey, Or sas^ey)
may be used to determine if the plume is derived from
natural or anthropogenic sources (e.g., Zielinski et al.,
1997). This information may be needed to determine the
site-specific capacity for uranium immobilization under
site-specific conditions. Also, as noted by Payne and Airey
(2006), determination of the ratios of these radioisotopes
in both ground water and aquifer solids can be used as a
direct indicator of active attenuation, where activity ratios
are anticipated to be similar for the two media where solid
phase uranium concentration is dominated by the uranium
mass from active attenuation within the aquifer. Both alpha
60
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spectrometry and plasma-based mass spectrometry can be
used to obtain these data (e.g., Cizdziel et al., 2005; Zheng
and Yamada, 2006; Stirling et al., 2007; Weyer et al., 2008).
Solid Phase Measurements
While it is evident that precipitation of uranium minerals
may occur under reducing or oxidizing conditions, direct
detection of these minerals via conventional methods such
as X-ray diffraction will not be feasible in most cases due
to low concentrations in aquifer solids. Use of microscopic
techniques based on electron scattering/diffraction (Buck
et al., 1996; Murakami et al., 1997; Jerden and Sinha, 2003;
Murakami et al., 2005) or X-ray absorption (Aral et al., 2007)
may be used for direct identification of uranium precipitates
in aquifer solids. However, these techniques are generally
labor-intensive and costly, and there are limitations to the
ability to analyze a statistically representative mass or
number of samples coincident with the dimensions of the
contaminant plume. As demonstrated by Catalano and
Brown (2004), accurate identification of individual solid
species in heterogeneous mixtures using X-ray absorp-
tion spectroscopy will be limited by the similarities in the
structures of uranium precipitates likely to form. For these
reasons, use of liquid extractions, either for the determina-
tion of the total concentration of uranium in aquifer solids
or the association of uranium with various components of
the aquifer solids, provide the most tenable approach to
ascertain the speciation or stability of attenuated uranium
(Amonette et al., 1994). Extraction-based methods may
be supplemented with characterization of select samples
using electron microscopy, TRLIFS (Wang et al., 2005a;
Wang et al., 2005b) or X-ray spectroscopy for confirmation
of speciation assignments. As an example, use of X-ray
absorption spectroscopy to determine the oxidation state
of solid phase uranium (Yamamoto et al., 2008) may be
valuable for confirming the attenuation mechanism.
Extraction of soils/sediments in combination with deter-
mination of specific radioisotope abundance in the liquid
extracts has been used to identify potential sources of
uranium contamination. Howe et al. (2002) used extrac-
tion solutions designed to target certain solid components
in a series of sediment samples collected as a function of
distance from a suspected contamination source to deter-
mine the influence of discharge from a nuclear fuel enrich-
ment plant. Total uranium concentrations in sediments
decreased with distance from the suspected source, and
the 238(j/235U ratjo measured were significantly lower than
those observed for a reference site with 238U and 235U at
natural abundance (238U/235U = 137.5). Oliver et al. (2008)
similarly used sequential extraction solutions in combination
with determination of isotopic abundance of extracted 238U
and 235U to assess the mobility of uranium derived from
depleted uranium in soils at weapons test ranges. In both
cases, it was possible to directly determine the contaminant
signature from that of natural sources of uranium. The use
of sequential extractions also provided indirect information
on the partitioning mechanism for contaminant sources of
uranium, as well as relative changes in the solid phase
speciation along the transport pathway. This approach is
limited by non-selective extraction of uranium associated
with specific solid components (Schultz et al., 1999) or
re-adsorption to un-extracted solid components (Schultz
et al., 1998; Lucey et al., 2007).
Determination of extractable U(VI) using bicarbonate solu-
tions has been evaluated by several research groups using
kinetic phosphorimetry as the detection method. Elias et al.
(2003) demonstrated that a 1 M NaHCO3 solution (pH = 8.3)
extracted freshly precipitated uranyl phosphate and uranyl
hydroxide spiked into aquifer sediment samples when con-
ducted under a nitrogen atmosphere. These authors also
demonstrated that this solution did not extract U(IV) from
reduced sediment samples under nitrogen, so this proce-
dure could be used to monitor the U(VI) content of solid
samples that may also contain U(IV)-bearing solids. Kohler
et al. (2004) evaluated use of a mixed solution of 0.014
M NaHCO3 and 0.0028 M Na2CO3 (pH = 9.45) to extract
U(VI) from contaminated aquifer solids that had been dried
in air. Additional evaluation of this procedure for contami-
nated sediment samples from this site indicated that the
extraction must be conducted under nitrogen immediately
after collection in order to avoid oxidation and extraction
of U(IV) that may be present in reduced sediments (Davis
et al., 2006). These authors again provided evidence that
U(IV) solid species are not extracted by sodium bicarbon-
ate solutions if the procedure is conducted under nitrogen
prior to potential oxidation of U(IV) during sample storage
or processing. While solutions of sodium bicarbonate
appear to dissolve both adsorbed and precipitated forms of
U(VI), this approach appears to provide a reliable measure
of U(VI) content in the contaminated aquifer solids. The
amount of U(IV) in the solids could then be determined
following more aggressive extraction using nitric acid (such
as in EPA SW-846 Method 3051) with detection of total
extracted uranium.
Long-Term Stability and Capacity
The long-term stability of attenuated uranium will depend
on the maintenance of either 1) ground-water chemistry
that prevents solubilization of U(VI) precipitates (e.g., phos-
phates or silicates), 2) sufficiently low reduction potentials
to prevent oxidation and consequent solubilization of U(IV)
solids, or 3) stability of the sorbent mineral and sufficiently
low concentrations of competing ions that could displace
the sorbed uranyl ion. Once uranium has been precipitated
or adsorbed, the sustainability of the geochemical driving
force (e.g., phosphate/silicate, redox, pH, and/or available
surface sites) is critical to whether natural attenuation will
be a viable cleanup option. Thus, it is recommended that
post-attenuation changes in water chemistry be carefully
considered to ensure that re-mobilization of attenuated ura-
nium does not occur. Of particular concern are situations in
which uranium is attenuated under reducing conditions that
are induced by characteristics of the contaminant plume,
specifically if the natural conditions within the aquifer are
more oxidizing. As reviewed by Suzuki and Suko (2006),
uranium concentrations in ground water may decrease to
acceptable levels as a result of uraninite precipitation under
reducing conditions. However, numerous studies have
61
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shown that oxidation and dissolution of the newly formed
uraninite occurs readily upon the onset of more oxidizing
conditions (Wan et al., 2005; Moon et al., 2007; Wu et al.,
2007; Komlos et al., 2008). Dissolved oxygen and nitrate
are common oxidants in ground water that can cause re-
mobilization of reduced uranium from aquifer solids.
Determination of capacity for attenuation will depend on
knowledge of the specific mechanisms leading to uranium
partitioning to aquifer solids and the flux of uranium being
transmitted through the aquifer. Since uranium may be
derived from both anthropogenic and natural sources,
determination of the source of this contaminant is critical
to assessment of available capacity for attenuation. For
anthropogenic sources of contamination, it is important
to understand the characteristics of uranium release
into ground water. The Hanford 300 Area in Richland,
Washington provides a clear example of how inadequate
characterization of the spatial- and time-dependent release
characteristics from uranium contaminant sources can
lead to an inaccurate assessment of available capacity for
attenuation within an aquifer (USEPA, 2008). The affective
capacity for attenuation within the aquifer will also depend
strongly on the characteristics and variability of ground-
water chemistry and aquifer solids properties along trans-
port pathways, as well as the impact of hydrologic dynamics
on subsurface chemistry as a function of space and time.
Tiered Analysis
Determination of the viability of uranium remediation in
ground water via monitored natural attenuation will depend
upon proper assessment of contaminant loading to the
aquifer and prevailing geochemistry and mineralogy within
the contaminant plume and the down gradient zone prior to
the point(s) of compliance. Due to the long half-lives for the
uranium radioisotopes of most concern, radioactive decay
will not provide a viable mechanism for plume attenuation.
Therefore, the goal of site assessment will be to demon-
strate the process(es) controlling uranium sequestration
onto aquifer solids and the long-term stability of solid phase
uranium as a function of existing and anticipated ground-
water chemistry. A recent technical review highlights several
technical aspects that need to be carefully evaluated at a
site in order to insure that reliable projections of attenuation
capacity and stability can be realized (USEPA, 2008). The
following tiered analysis structure for site characterization
provides a technically defensible approach to evaluate
candidate sites and define the potential limitations of MNA
as part of a remedy for ground-water cleanup.
Tier I. Site characterization under Tier I will involve demon-
stration that the ground-water plume is static or shrinking,
has not reached compliance boundaries, and does not
impact existing water supplies. Once this is established
through ground-water characterization, evidence is col-
lected to demonstrate uranium partitioning to aquifer solids
within the plume. If natural attenuation processes are active
throughout the plume, then there should be an observed
increase in solid phase concentrations within regions
of the plume with higher aqueous concentrations, e.g.,
near the source term. This field partitioning data may be
supplemented by geochemical modeling that incorporates
measured water chemistry (e.g., pH, Eh, and major ion
chemistry) throughout the plume to assess the potential
for solubility control by a uranium precipitate such as an
oxide, silicate, or phosphate phase. Identification of active
sequestration to prevent uranium migration in ground-water
provides justification for proceeding to Tier II characteriza-
tion efforts.
Tier II. Under Tier II, the apparent rate and mechanism(s)
of attenuation are determined. Estimates of a site attenu-
ation rate(s) can be assessed via a well transect along
the ground-water flow path. In addition, time-series data
may be collected at one or more monitoring points within
the plume. This information will allow assessment of the
relative timescales for contaminant immobilization and
fluid transport and determination of whether remediation
objectives can be met within the required regulatory time
frame. In addition, the mechanism(s) for attenuation are to
be identified under this stage of site characterization. This
effort will require determination of the chemical speciation of
aqueous and solid phase uranium and may be approached
according to the following scheme:
1) Determination of solution speciation via direct
analytical measurements to define dissolved
uranium oxidation state and aqueous complexation
(e.g., Sowder et al., 1998; Billard et al., 2003) in
combination with speciation calculations based on
characterized ground-water chemistry;
2) Determination of the oxidation state of solid phase
uranium (e.g., Elias et al., 2003);
3) Calculation of saturation state of ground
water relative to measured aqueous chemistry
complimented by the possible isolation of discrete
uranium mineral phases via density separations
(or other schemes) in regions of the aquifer with
highest solid phase concentrations;
4) Determination of aquifer mineralogy (Amonette,
2002) to determine the relative abundance of
components with documented capacity for uranium
sorption (e.g., Davis et al., 2004); and
5) Determination of uranium-sediment associations
via chemical extractions designed to target specific
components within the aquifer sediment (e.g.,
Schultz et al., 1998; Oliver et al., 2008).
This compilation of information will facilitate identification
of the reaction(s) leading to uranium immobilization. It is
recommended that identification of uranium chemical spe-
ciation in aqueous and solid matrices be conducted using
samples collected in a manner that preserves the in-situ
speciation of dissolved uranium and mineralogy (Davis
et al., 2006; USEPA, 2006b) and prevents loss of uranium
from aqueous samples (e.g., due to oxidation and precipita-
tion of ferrous iron in anoxic ground water). The demonstra-
tion of concurrence between conceptual and mathematical
models describing uranium transport will entail development
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of site-specific parameterization of the chemical processes
controlling uranium solid phase partitioning.
Tier III. Once the partitioning mechanism(s) have been
identified for the site, the subsequent characterization effort
under Tier III will involve determination of the stability of
immobilized uranium and the capacity of the aquifer to sus-
tain continued uptake. It is recommended that the stability
of immobilized uranium be tested based on the anticipated
evolution of ground-water chemistry concurrent with decay
of the plume. For example, changes in ground-water pH
can exert a significant influence on uranium adsorption or
precipitate solubility. Therefore, it is recommended that
sediment leach tests be conducted to characterize the
magnitude of uranium mobilization as a function of pH for
a ground-water chemistry representative of site conditions.
It is recommended that the capacity for uranium uptake
onto aquifer solids be determined relative to the specific
mechanism(s) identified in Tier II. For example, if site
characterization under Tier II indicated that microbial deg-
radation of naturally occurring solid organic matter (SOM)
resulted in reduction of U(VI) to insoluble U(IV), then the
mass distribution of SOM within the aquifer needs to be
determined. This site-specific capacity would then be com-
pared to uranium mass loading within the plume in order
to assess the longevity of the natural attenuation process.
Evaluation of uranium radioisotope distributions in samples
of ground water and aquifer solids along transport pathways
is recommended in order to confirm whether the source of
uranium is from identified contaminant source areas or from
natural sources that are not stable under plume chemical
conditions. If site-specific tests demonstrate the stability
of immobilized uranium and that there is sufficient capac-
ity within the aquifer to sustain uranium attenuation, then
the site characterization effort can progress to Tier IV. For
cases where contaminant stability is sufficient but aquifer
capacity is insufficient for capture of the entire plume, then
a determination of the benefits of contaminant source
reduction is required.
Tier IV. Finally, under Tier IV a monitoring plan is estab-
lished along with contingency plans in the event of MNA
failure. It is recommended that the monitoring plan be
designed to establish both continued plume stability and
to identify changes in ground-water chemistry that may
lead to re-mobilization of attenuated uranium. The specific
chemical parameters to be monitored will include those
identified under Tier III that may halt uranium partitioning
and/or result in dissolution of either discrete uranium pre-
cipitates or aquifer minerals that sequester uranium from
ground water. For example, solution phase parameters
that could alter either uranium precipitation or adsorption
include inorganic carbon (alkalinity), major ion chemistry
such as Ca/Mg, and/or pH. In contrast, the concentration of
dissolved iron may indicate the dissolution of an important
sorbent phase within the aquifer (e.g., reductive dissolution
of iron oxides). Changes in these parameters may occur
prior to observed changes in solution uranium and, thus,
serve as monitoring triggers for potential MNA failure. In
this instance, a contingency plan can be implemented
that incorporates alternative strategies to arrest possible
plume expansion beyond compliance boundaries. Possible
strategies to prevent plume expansion include pump and
treat operations, installation of reactive barriers to enhance
uptake capacity perpendicular to the direction of plume
advance, or enhancement of attenuation processes within
the aquifer through the injection of soluble reactive com-
ponents (e.g., injection of phosphate to drive precipitation
of autinite-like phases; See Nimmons, 2007 and USEPA,
2007 for example technologies.).
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Iodine
Daniel I. Kaplan, Robert G. Ford, Richard T. Wilkin
Occurrence and Distribution
Iodine occurs as a trace element in the Earth's crust (aver-
age abundance of 0.45 mg/kg), and the concentration
of iodine in soil ranges from 0.5 to 40 mg/kg with higher
values commonly associated with organic soils (Fuge and
Johnson, 1986; Hou et al., 2009). There are 14 radio-
isotopes of iodine with half-lives greater than 10 minutes
(Hou et al., 2009). The only stable isotope of iodine is
127I and the most long-lived radioisotope is 129I. The 129I
radioisotope is produced in nature by spontaneous fission
of naturally-occurring uranium (spontaneous fission of 238U
and thermal-neutron induced fission of 235U in the crust) or
by the interaction of cosmic ray particles with xenon in the
atmosphere (Schmidt et al., 1998; Hou, 2004), although
the 129|/127| ratio from natural sources is substantially lower
than that from human nuclear activity (Hou et al., 2009).
Common anthropogenic sources of 129I include the produc-
tion and reprocessing of nuclear fuel, since 129I is produced
as a fission product with a fission yield of 0.9% from 235U
and 1.6% from 239Pu (USEPA, 2004; Aldahan et al., 2007).
Nuclear-fuel reprocessing facilities constitute the major
source of 129I released in the environment. For example, it
has been estimated that two European facilities at La Hague
in France and Sellafield in England released 2,300 kg of
129I to the environment from the start of operations through
1998 (Hou et al., 2000; Frechou and Calmet, 2003). For an
example within the United States, between 1955 and 1989,
the Savannah River Site in South Carolina released 52 kg
129I (Kantelo et al., 1993), a portion of which was released
to the subsurface via seepage basins (Beals and Hayes,
1995; Denham et al., 2009). Release of 129I into the sub-
surface has also been documented for the Idaho National
Laboratory (e.g., Beasley et al., 1998; Cecil et al., 2003),
and it is suggested that historical nuclear fuel reprocessing
activities have served as a primary source of anthropogenic
129I across North America (Rao and Fenn, 1999).
Geochemistry and Attenuation Processes
Radioactive Decay
Iodine has one stable isotope in nature, 127I. Radioactive
129I, with a half-life of 1.57x1O7 years, is an important fission
product with a fission yield of 0.9% from 235U and 1.6%
from 239Pu. Fission of 235U also leads to the production
of 131I. However, due to its short half-life of 8.04 days, 131I
releases do not typically generate ground-water plumes of
significant extent.
Aqueous Speciation
In natural systems, iodine may exist in the -1, 0, +1, +5, and
+7 oxidation states (e.g., Burger and Liebhafsky, 1973; Hou
et al., 2009). Of these, the -1 (e.g., aqueous h or "iodide")
and +5 (e.g., aqueous IO3 or "iodate") oxidation states are
the most abundant inorganic species in ground water. The
stability range of I" encompasses almost the entire pH and
Eh range typically encountered in ground water (Figure 7.1).
Under acidic conditions (pH<4), L . (oxidation state 0;
"iodine") can be produced from the oxidation of I" or the
reduction of IO3". Iodine partitions between a soluble spe-
cies [I2(aq)] and a dissolved gas [I2(g)] with the volatile form
being susceptible to transfer into air (Evans et al., 1993).
Iodide can form complexes with metal ions, but these are
generally the least stable of all the halide complexes, with
a few notable exceptions. Iodide forms extremely strong
complexes to some soft metals, including silver and mer-
cury (Gammons and Yu, 1997; Pruszyriski et al., 2006).
Iodate can also form complexes with a range of metal ions
in aqueous systems (e.g., Miyamoto et al., 2008). Iodine
forms relatively strong chemical bonds with organic matter,
where iodine covalently bonds with carbon in the molecular
structure of natural organic matter compounds (Walters and
Winchester 1971; Schlegel et al., 2006). Several terms
are used in reference to organic iodine species, including
"iodoorganic compounds", "iodinated organic compounds",
and specific forms such as methyl iodide (e.g., Amachi et
al., 2001; Schwehr et al., 2009). Aqueous I2 and methyl
iodide may be transported from ground-water systems as
volatile species into the overlying vadose zone or shallow
soils (Fuge and Johnson, 1986; Fuge, 1990). As discussed
below, the formation of volatile iodine species may be medi-
ated by abiotic or biotic reactions.
The redox speciation of iodine in ground water may be
influenced by a combination of abiotic and biotic reactions.
Abiotic reactions that have been documented to result in a
change in the oxidation state of iodine include reaction with
Fe- or Mn-bearing constituents, as well as natural organic
matter compounds. For reaction with Fe species, reduction
of IO3~ to I" by ferrous iron or Fe(l l)-bearing solid phases has
been observed in controlled laboratory studies (Jia-Zhong
and Whitfield, 1986; Councell et al., 1997; Hu et al., 2005;
Glaus et al., 2008). These reactions may be influenced
by the activity of iron-reducing bacteria stimulated within a
plume with degradable organic co-contaminants. Natural
organic matter (e.g., fulvic acid) and Mn(ll) also have the
capacity to reduce IO3" and/or I2(aq) to I" (e.g., Skogerboe
and Wilson, 1981; Anschutz et al., 2000; Steinberg et al.,
2008a). The reduction of IO3" by natural organic matter may
ultimately lead to incorporation of iodine into the organic
matter molecule via formation of hypoiodous acid (HIO) or
L , as an intermediate (Steinberg et al., 2008a). In general,
these reactions are viewed as an electrophilic substitution
of hydrogen by iodine on a phenolic ring within the natural
organic matter structure (Reiller et al., 2006). Iodine forms
a covalent bond with the phenolic group, and this reaction
69
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may be preceded by the reduction of IO3 (e.g., Steinberg
et al., 2008a). Iodide (I") may also be oxidized to IO3" dur-
ing interaction with Mn(IV) oxides present in aquifer solids
(Fox et al., 2009; Gallard et al., 2009). In the presence of
natural organic matter, Mn(IV) oxide may catalyze formation
of organoiodide compounds (Gallard et al., 2009; Schwehr
etal.,2009).
LLJ
-.5
25°C
4 5 6 7 8 9 10
PH
Figure 7.1. Eh-pH diagram of dominant iodine aqueous
species at 25°C (total I = ia6 mol/L).
Methylation and Volatilization
Sheppard et al. (2006) summarize available literature on
the rate of volatile loss of iodine from soils. In general,
review of available information for terrestrial systems
indicates that the half-life for volatilization from soils is on
the order of a decade or more. The potential for volatile
losses will be mediated by characteristics of soils overlying
the iodine plume, where soils containing higher content of
organic carbon can act as a sorptive barrier for migration of
volatile iodine species (Bostock et al., 2003). In this case,
the effectiveness of the soil organic carbon for sorption of
volatile iodine species may be limited by the form of volatile
iodine, with I2(aq) being the most susceptible for sorption
(see discussion above). Microbially-mediated production
of volatile species of iodine appears to lead to formation of
methyl iodide (CH3I) as the primary volatile form (Amachi
et al., 2001; Amachi et al., 2003; Muramatsu et al., 2004).
While Muramatsa and Yoshida (1995) observed higher
volatile losses of iodine under anoxic conditions where
microbially-mediated production of methyl iodide appeared
to dominate, the primary conduit for methyl iodide emis-
sions was through plant shoots. Thus, it appears that direct
volatile losses of iodine in shallow groundwater systems
will be minimal, consistent with current recommendations
discussed in Sheppard et al. (2006).
Adsorption
The adsorption of iodine onto various minerals that may
be present in aquifer solids has been reviewed in USEPA
(2004). For systems in which I" and IO3" are the predomi-
nant species in groundwater, it is generally observed that
IO3" adsorbs to a greater extent (Fukui et al., 1996; Hu et al.,
2005; Kodama et al., 2006). The extent of 129I adsorption
will be limited by the presence of naturally occurring 127I that
may also be present in ground water and/or partitioned to
aquifer solids (e.g., Tournassat et al., 2007). Another impor-
tant factor for iodine adsorption is partitioning to natural
organic matter (NOM) in aquifer solids. There are several
studies that document the partitioning of iodine to organic
carbon in aquifer solids via covalent bonding (Schlegel
et al., 2006; Steinberg et al., 2008b), where I2(aq. appears
to be a necessary intermediate. As discussed above, the
intermediate I2(aq) species might be formed at low pH or via
the abiotic/biotic reactions that lead to the oxidation of I",
whereby iodine may be incorporated into organic species
upon reaction with NOM (e.g., Warner et al., 2000; Reiller
et al., 2006; Yamaguchi et al., 2006). As demonstrated by
Gallard et al. (2009) and Schwehr et al. (2009), the rela-
tive distribution of iodine species resulting from the various
redox reactions highlighted above will depend on the total
iodine concentration as well as the relative abundance of
reactants such as NOM and Mn(IV) oxides.
The hydrodynamics of the ground-water system may also
exert control on the adsorption of iodine by influencing the
total flux of oxidants or reductants that react with iodine
during solid phase partitioning. This behavior is illustrated
in the study by Ashworth and Shaw (2006a) where the
dynamics of iodine solid-phase partitioning was monitored
as a function of water content and redox potential. These
dynamics are similar to what might be encountered at a
fluctuating ground-water table at the "smear" zone that is
commonly established at the interface between saturated
and unsaturated aquifer solids. Ashworth et al. (2003;
2006b) have demonstrated this behavior in lysimeter experi-
ments conducted in a manner to replicate the oxic-anoxic
zonation that may develop at the water table in shallow
ground-water systems. These studies have demonstrated
accumulation of total iodine within the oxic-anoxic transition
zone, which is consistent with observations of the vertical
distribution of stable iodine in subsurface systems (e.g.,
Yuita and Kihou, 2005).
Site Characterization
Overview
Attenuation of 129I might be achieved through sorption to
aquifer solids. In general, two factors that appear to exert
greatest influence on uptake of iodine onto aquifer solids
are the predominance of the IO3" species and/or the pres-
ence of immobile NOM that can bind to iodine for systems
where production of I2(aq) is active. These mechanisms will
be influenced by the subsurface redox chemistry, which
may be poised by available oxidants/reductants in ground
water and aquifer solids or microbial processes that may be
70
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active throughout the plume. A list of potential attenuation
processes is provided in Table 7.1.
For some sites, the volatile transfer of L , or methylated
species of 129I from shallow ground water into overlying
unsaturated soil or directly into open or confined atmo-
spheres may also reduce ground-water concentrations.
However, reliance on this mass transfer process will neces-
sitate determination of the potential consequence of this
exposure pathway to human and/or ecosystem health.
Unacceptable exposures to airborne 129I via volatile loss
from ground water may necessitate active control of 129I
mass transport through the aquifer.
Table 7.1 Natural attenuation and mobilization path-
ways for iodine.
Attenuation
Processes
Adsorption of
IO3~ or l~ onto
aquifer minerals
(clay minerals,
iron/manganese
oxyhydroxides)
Covalent
bonding to
immobile forms
of NOM in
aquifer solids
Mobilization
Processes
Desorption
due to
increasing pH
or competition
from major
anions in
ground water.
Degradation or
mineralization
of NOM
resulting in
release of
complexed
iodine;
dissolution/
desorption
of NOM into
ground water
Characterization
Approach
Evaluate total adsorption
capacity of aquifer solids
under representative
ground-water chemistry;
chemical extractions to
assess concentrations
and speciation of
129I sorbed to aquifer
solids along relevant
transport pathways and
determination of mobile
iodine species in plume
(127I and 129I).
Determine concentration
of aqueous and solid
forms of NOM and
bound 129I along
transport pathway(s).
Aqueous Measurements
Since 129I decays by emitting p-particle with a maximum
energy of 154.4 keV and y-rays of 39.6 keV as well as X-rays
(29-30 keV), it can be measured using y-spectrometry
and p-counting using liquid scintillation counting (Hou
and Roos, 2008; Hou et al., 2009). The presence of other
radionuclides with overlapping decay energy emissions
(e.g., 137Cs gamma peak at 32.1 keV) typically necessitates
prior chemical separation of 129I from the sample matrix.
Alternatively, inductively coupled plasma-mass spectrom-
etry (ICP-MS) may also be used for the quantification of 129I
via direct introduction of aqueous solutions or volatilization
of iodine and introduction of the volatile species (Wuilloud
and Altamirano, 2006; Brown et al., 2005; Wang and Jiang,
2008; Li et al., 2009). Recent developments in sample
introduction schemes and techniques to minimize interfer-
ence from formation of molecular ions within the plasma
such as 127I1H1H+ or interference from ions such as 129Xe+
has improved the ability of ICP-MS for detection of 129I at
regulatory levels (Grinberg and Sturgeon, 2009). However,
it should be noted that determination of 129I by ICP-MS may
ultimately be limited by the relative abundance of 127I, with a
reported lower limiting value of the 129|/127| mass concentra-
tion ratio of 10~7 in the analyzed sample (Li et al., 2009).
For the purpose of analyzing both total iodine and individual
species of iodine in ground water, aqueous samples should
not be acidified and are best stored in a gas-tight sample
bottle with no head space. Acidification is to be avoided
due to the potential pH-dependent species interconversions,
e.g., reaction of I" and IO3" to form I2 (e.g., Cripps et al.,
2003; Bhagat et al., 2008; Reid et al., 2008). It should
also be noted that acidification of unfiltered ground-water
samples may also be a source of high bias in measured
dissolved iodine concentrations due to release from particu-
lates in the sample (e.g., Buraglio et al., 2000). It is recom-
mended that this situation be recognized and avoided in
the monitoring system, since the resultant data may not be
useful for evaluating mechanisms controlling contaminant
transport. Tagami and Uchida (2005) investigated potential
losses of total iodine from natural water samples for several
storage conditions and holding times. In general, it has
been observed that alkaline pH adjustment and storage
below 5°C minimized analyte loss during storage (Tagami
and Uchida, 2005; Wei et al., 2007). Losses of volatile spe-
cies during sampling and handling could be minimized via
direct sampling using in-situ diffusive membrane samplers
deployed within the sampled water body or in-line during
pumping from the ground-water well (e.g., Groszko and
Moore, 1998).
Speciation of various iodine species in water samples
can be conducted using chromatographic (Schwehr and
Santschi, 2003; Yang et al., 2007; Wang and Jiang, 2008;
Hou et al., 2009) or non-chromatographic (Bruchertseifer
et al., 2003; Gonzalvez et al., 2009) methods of separation.
For methods in which species separation and analysis is
conducted in the laboratory following sample collection,
confirmation of adequate preservation techniques is needed
to demonstrate maintenance of in-situ chemical specia-
tion of iodine. This can be accomplished by inclusion of
field matrix spikes with additions of known iodine species
at the point of sample collection. Field separation of I"
and IO3~species may be achieved through use of anion-
exchange resins (e.g., Bruchertseifer et al., 2003; Cripps
et al., 2003) with the potential to capture dissolved forms
of potentially volatile species such as I2(aq) or CH3I using
solid-phase extraction (Bruchertseifer et al., 2003). Amachi
et al. (2000) have proposed an alternative approach to
isolate volatile inorganic and organic species of iodine from
aqueous samples based on purge-and-trap in a container
head space. As recommended in the review by Hou et al.
(2009), the time between sampling and separation of iodine
species should be minimized and is best conducted at the
point of sample collection.
71
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Solid Phase Measurements
Several methods have been documented for the analysis
of total iodine in solid samples. These methods can be
generally grouped as those that employ extraction of iodine
from the solid matrix with alkali in liquid or solid form (Bing
et al., 2004; Brown et al., 2005; Mani et al., 2007) or those
that employ thermal extraction of iodine under an oxygen
atmosphere (e.g., Izmer et al., 2003; Izmer et al., 2004; Chai
et al., 2007, Englund et al., 2007; Balcone-Boissard et al.,
2009). For the thermal extraction methods, extracted iodine
is either trapped in an alkaline solution prior to analysis or
delivered to the detection system in gas phase via direct
coupling. Mani et al. (2007) describe use of a solution of
tetra methyl ammonium hydroxide (TMAH) under relatively
mild conditions for extraction of iodine from reference soils,
sediments, and geochemical exploration samples. The
values determined using extraction with TMAH compared
well with published certified values for the reference soils.
As has been demonstrated by Szidat et al. (2000a), it is
important to assess the potential for sample contamina-
tion due to carryover throughout the analytical system.
Assessment of carryover or losses of iodine species needs
to include all processes of the entire analysis, including
sampling, storage and sample preparation (Hou et al.,
1998; Szidat etal.,2000b).
Distinguishing between stable (127I) and contaminant forms
of iodine (129I) partitioned to aquifer solids may be neces-
sary to properly identify the process(es) resulting in plume
attenuation. Several published studies have made use of
radiometric and mass-specific detection methods in order
to identify the mass distribution of iodine isotopes, includ-
ing evaluation of individual isotopes in reference soil/sedi-
ment materials. Values for 127I concentration in Standard
Reference Material 2709 (San Joaquin Soil) has been
determined via several procedures involving either direct
thermal volatilization of iodine from the solid matrix (e.g.,
Resano et al., 2005; 4.9 ± 0.3 mg/kg), thermal combus-
tion and liquid trapping of 127I (e.g., Marchetti et al., 1997;
4.67 ± 0.32 mg/kg), or decomposition of the soil via alkaline
fusion followed by dissolution into a liquid matrix (Brown
et al., 2005; 3.77-5.21 mg/kg). Marchetti et al. (1997) also
report 127I concentrations for Standard Reference Materials
2711 (Montana Soil; 2.67 ± 0.9 mg/kg) and 2704 (Buffalo
River Sediment; 1.76 ± 0.07 mg/kg). Applications for the
analysis of 129I in solid materials have also been docu-
mented. For example, an updated measurement for the
concentration of 129I in the IAEA-375 reference soil mate-
rial has been reported (Jiang et al., 2005). Schmidt et al.
(1998) examined the influence of different approaches to
extract and to assess recovery of 129I and 127I recovered
from IAEA-375. Roberts and Caffee (2000) reported the
results of an interlaboratory comparison study to assess
the variability in 129I concentrations determined for solid
samples, which indicated that relative reporting errors could
be high without adequate quality assurance measures. In
addition, Izmer et al. (2004) have reported measurement
of the 129|/127| ratio for Standard Reference Material 4357
(Ocean Sediment; 5.3x10'7 versus the certified 4.45x10~7)
using direct analysis of vapor by ICP-MS following extrac-
tion of iodine from the sediment by heating in an oxygen
atmosphere. However, as discussed by Hou and Roos
(2008), reliable determination of r29/P27 ratios <10x-10 (i.e.,
a pre-nuclear age ratio) may require the use of accelerator
mass spectrometry. Analysis of the 129|/127| ratio, in addition
to the total 129I concentration, provides a means to demon-
strate 129I plume attenuation onto subsurface solids along
relevant ground-water flow paths.
Identification of the attenuation mechanism(s) that control
iodine sorption within the plume may also necessitate
determination of the chemical form of solid-phase iodine.
While there are several approaches to chemically extract
iodine from solid matrices for determination of total con-
centration, there are not established methods to extract
iodine in a manner that preserves information on the nature
of the chemical form of iodine originally in the solid matrix.
Associations of iodine with specific solid components, e.g.,
iron oxyhydroxides or organic carbon, are often inferred
based on apparent co-extraction of iodine with elements of
the dissolved solid component (e.g., Fitoussi and Raisbeck,
2007; Hou et al., 2009 and references therein). As an
example, Fitoussi and Raisbeck (2007) employed extraction
of marine sediments with an alkaline TMAH solution in an
effort to target 129I associated with organic matter. However,
further development of these types of extraction procedures
is needed to verify the selectivity of the co-extraction pro-
cess, since aggressive chemical solutions may also extract
129I associated with other solid components. Alternatively,
more recent studies have made use of X-ray absorption
spectroscopy (XAS) to determine the in-situ oxidation state
of solid-phase iodine and discern solid phase associations
by reference to XAS spectra in model compounds (Kodama
et al., 2006; Schlegel et al., 2006; Yamaguchi et al., 2006;
Shimamoto and Takahashi, 2008). This technique provides
the ability to differentiate between inorganic and organic
forms of solid-phase iodine. While this analytical method is
not routinely available, it does provide an approach to avoid
potential analytical artifacts such as the oxidation/reduction
of iodine that may occur during chemical extraction.
Long-term Stability and Capacity
The relative capacity for iodine sorption to aquifer solids,
as well as the relative stability of sequestered 129I, will likely
depend on the effective period of contact between ground
water and aquifer solids. As shown by Kaplan et al. (2000)
and Urn et al. (2004), systems in which there is a relatively
short period of equilibration between 129I and aquifer solids
may be more susceptible to remobilization. As discussed
above, the stability of adsorbed inorganic iodine species
may also be influenced by changes in system redox, where
the onset of reducing conditions tends to result in desorp-
tion of inorganic iodine as the I" species. The formation of
covalent bonds between iodine and immobile organic matter
in aquifer solids is not a fast process, so this attenuation
process may be limited for systems with higher ground-
water flow velocities. The speciation of iodine in ground
water and the sorption characteristics of aquifer solids will
influence the relative importance of ground-water velocity
72
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on the degree of 129I attenuation. Relative to the sorption
characteristics of aquifer solids, Alvarado-Quiroz et al.
(2002) have presented field data demonstrating changes
in the apparent partitioning behavior of 129I and 127I along
a ground-water flow path that transitions from surficial
sands into a wetland area with elevated levels of sphagnum
peat. The apparent in-situ partitioning constant for iodine
measured along this flow path varied by 2-3 orders-of-
magnitude, increasing with the change to a peaty texture
in aquifer solids at the point of plume discharge within the
wetland. As shown by Schwehr et al. (2005) for a coastal
aquifer in California, attenuation of iodine may not be sig-
nificant in situations where relatively high concentrations
of competing anions exist (e.g., Cl~) and the mobile form
of 129I is dominated by the iodide chemical species (I").
The capacity for sorption of 129I will also be limited by the
concentration of stable iodine (127I) within the plume. For
example, in the F-Area plume on the Savannah River Site
(Kantelo et al., 1993), background concentrations of stable
iodine were 600 times greater than the 129I concentrations
(stable iodine = 270 ug/L and 129I = 0.45 ug/L or 80 pCi/L).
Relative to iodine release from uncontrolled source areas,
knowledge of the total contaminant mass as well as the rate
and frequency of release into the saturated aquifer needs
to be developed. In order to make a reliable assessment
of the mass/activity flux of iodine into the plume, it will be
important to understand the characteristics of the hydrogeo-
logic system and the dynamics of water and contaminant
transfer from contaminant source areas into the plume.
Tiered Analysis
Determination of the viability of 129I remediation in ground
water via monitored natural attenuation will depend upon
proper assessment of contaminant loading to the aquifer
relative to the sorption capacity of aquifer solids along
relevant transport pathways to the point(s) of compliance.
The goal of site assessment is to demonstrate that sorp-
tion is adequate to meet cleanup goals given current and
projected hydrologic conditions for the site. The following
tiered analysis structure for site characterization provides
an approach to evaluate candidate sites and define the
potential limitations of MNA as part of a remedy for ground-
water cleanup.
Tier I. Site characterization under Tier I will involve demon-
stration that the ground-water plume is static or shrinking,
has not reached compliance boundaries, and does not
impact existing water supplies. Once this is established
through ground-water characterization, evidence is col-
lected to demonstrate 129I partitioning to aquifer solids
within the plume. If natural attenuation processes are active
throughout the plume, then there should be an observed
increase in solid phase concentrations concurrent with
decreases in aqueous contaminant concentrations. This
field partitioning data may be supplemented by geochemi-
cal modeling that incorporates measured water chemistry
(e.g., pH, Eh, and major ion chemistry) throughout the
plume to assess the predominant oxidation state of iodine
within the plume. Identification of active sequestration to
prevent 129I migration in ground-water provides justification
for proceeding to Tier II characterization efforts.
Tier II. Under Tier II, the apparent rate and mechanism(s)
of attenuation are determined. Estimates of a site attenu-
ation rate(s) can be assessed via a well transect along
the ground-water flow path. In addition, time-series data
may be collected at one or more monitoring points within
the plume (USERA, 2007; Section IIIA.5). This informa-
tion will allow assessment of the relative timescales for
contaminant immobilization and fluid transport and deter-
mination of whether remediation objectives can be met
within the required regulatory time frame. In addition, the
mechanism(s) for attenuation are to be identified under
this stage of site characterization. This effort will require
determination of the chemical speciation of solid phase 129I
and may be approached according to the following scheme:
1) Determination of solution and solid phase 129I
concentrations, along with the relative concentration
of major ions/components in aquifer solids where
attenuation is occurring, including analysis of
trends in the distribution of 129I with solid phase
components that may be representative of potential
sorbents;
2) Determination of aquifer mineralogy (Amonette,
2002) to determine the relative abundance of
components that might support iodine sorption,
e.g., iron oxyhydroxides and/or NOM; and
3) Determination of 129l-solids associations via
chemical extractions designed to target specific
components within the aquifer solids.
This compilation of information will facilitate identification of
the reaction(s) leading to 129I immobilization. The demon-
stration of concurrence between conceptual and mathemati-
cal models describing 129I transport will entail development
of site-specific parameterization of the chemical processes
controlling 129I solid phase partitioning.
Tier III. Once the partitioning mechanism(s) have been
identified for the site, the subsequent characterization effort
under Tier III will involve determination of the stability of
immobilized 129I and the capacity of the aquifer to sustain
continued uptake. It is recommended that the stability of
immobilized 129I be tested based on the anticipated evolu-
tion of ground-water chemistry concurrent with decay of
the plume. For example, changes in the concentrations of
dissolved halogen anions such as chloride or bromide can
result in desorption of iodine from aquifer solids. Therefore,
it is recommended that sediment leach tests be conducted
to characterize the magnitude of 129I mobilization as a func-
tion of halogen anion concentrations for a ground-water
chemistry representative of site conditions (e.g., Kaplan
et al., 2000). It is recommended that the capacity for 129I
uptake onto aquifer solids be determined relative to the
specific mechanism(s) identified in Tier II. For example, if
sight characterization under Tier II indicated that immobile
NOM was the primary sorbent leading to 129I attenuation,
then the mass distribution of NOM within the aquifer needs
to be determined (e.g., Alvarado-Quiroz et al., 2002). This
73
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site-specific capacity would then be compared to 129I mass
loading within the plume in order to assess the longevity of
the natural attenuation process. The flux or mass loading
of 129I through the aquifer source release characteristics
may be influenced by the physical location of sources
within the subsurface relative to the ground-water table.
Fluctuations in infiltration through shallow, unsaturated
zones and/or water table elevations within the aquifer due
to variations in recharge may lead to periodic increases
in 129I flux into ground water from contaminant source
areas. It is recommended that additional iodine transport
modeling be included to evaluate the impact of these vari-
ous scenarios to be assured that these perturbations do
not significantly diminish attenuation. If site-specific tests
demonstrate the stability of immobilized 129I and that there
is sufficient capacity within the aquifer to sustain 129I attenu-
ation, then the site characterization effort can progress to
Tier IV. For cases where contaminant stability is sufficient
but aquifer capacity is insufficient for capture of the entire
plume, then a determination of the benefits of contaminant
source reduction is recommended.
Tier IV. Finally, under Tier IV a monitoring plan is estab-
lished along with contingency plans in the event of MNA
failure. It is recommended that the monitoring plan be
designed to establish both continued plume stability and to
identify changes in ground-water chemistry that may lead
to re-mobilization of attenuated 129I. The specific chemical
parameters to be monitored will include those identified
under Tier III that may halt 129I partitioning and/or result
in dissolution of aquifer minerals that sequester "Tc from
ground water. For example, a solution phase condition
that could result in 129I desorption includes development
of reducing conditions within a previously oxic zone where
IO3" was the predominant attenuated species. In contrast,
increases in the concentration of dissolved ferrous iron
may indicate the dissolution of an important sorbent phase
within the aquifer (e.g., reductive dissolution of iron oxy-
hydroxides). For sites at which residual, subsurface 129I
sources are left near the water table, it is recommended
that the site monitoring program include locations to assess
changes in the release of 129I to the saturated aquifer due
to increased surface infiltration or rises in the ground-water
table. Changes in these chemical and hydrologic param-
eters may occur prior to observed changes in solution 129I,
and thus serve as monitoring triggers for potential MNA
failure. In this instance, a contingency plan can be imple-
mented that incorporates alternative strategies to arrest
possible plume expansion beyond compliance boundaries.
Possible strategies to prevent plume expansion include
pump and treat operations, installation of reactive barriers
to enhance uptake capacity perpendicular to the direction of
plume advance, or enhancement of attenuation processes
within the aquifer through the injection of soluble reactive
components that induce more oxidizing conditions.
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Radium
Daniel I. Kaplan, Robert G. Ford, Richard T. Wilkin
Occurrence and Distribution
Radium in ground water may be derived from both natural
and anthropogenic sources. Four radioisotopes of radium,
228Ra, 224Ra, 226Ra and 223Ra, are derived as decay products
from naturally-occurring 232Th, 235U or 238U radioisotopes
in soils and rocks. As such, occurrence and distribution
of radium in the natural environmental is largely con-
trolled by the distribution of thorium and uranium. Bowen
(1979) reported radium concentrations in rocks to range
from 0.6 x 10'6 to 1.1 x 10'6 mg/kg and in sediments to be
~0.8x 10'6 mg/kg. Radium concentrations in igneous rocks
increase by 100 times along the geological sequence: ultra-
basic (0.009 x 10'6 mg/kg), basic igneous (0.6 x 10'6 mg/kg),
intermediate igneous (0.917 x 10'6 mg/kg), and granitic
igneous (1.40 x 10'6 mg/kg) (Ames and Rai, 1978). The
radium content of granitic igneous, sandstones, shales and
limestones are all about the same, ~1 x 10'6 mg/kg.
There have been several reports relating geological land
forms and land utilizations to total radium concentrations
in drinking water (Rama and Moore, 1984; Longtin, 1988;
Michel and Cothern, 1986; Michel, 1990; dePaul and Szabo,
2007; Vinson et al., 2009). Geological formations contain-
ing elevated uranium or thorium concentrations have been
shown to contain elevated dissolved radium concentrations
in their porewaters (Michel, 1990). Ground water impacted
by agriculture has also been shown to have elevated radium
concentration (Sidle et al., 2001). Waste solids and prod-
ucts of phosphate fertilizer production may contain elevated
levels of uranium, and, subsequently, serve as a source of
radioisotopes of radium that could leach into ground water
(Rutherford etal., 1995; Rutherford et al., 1996; Kim etal.,
2006). Finally, extensive surveys of ground water wells
have shown that land uses that result in the acidification
of ground water, to pH levels less than 5, were correlated
to elevated radium concentrations (Michel and Cothern,
1986; Szabo and dePaul, 1998).
Anthropogenic sources of radium include phosphate mining
areas (uranium exists as impurities in rock phosphates at
concentrations of 100 to 150 mg/kg; Altschuler, 1973), ura-
nium mining areas (Landa and Gray, 1995), areas impacted
by oil-field brines (Pardue and Guo, 1998; Zielinski and
Budhan, 2007; Bou-Rabee et al., 2009), and facilities where
nuclear materials have been manufactured or processed
(Siegel and Bryan, 2003). Other potential localized sources
of radium include ground-water well drilling fluids whose
composition includes natural barite with varying levels of
coprecipitated radium (Clark et al., 2004), as well as lumi-
nescent devices that historically made use of radium as a
material component (Baker and Toque, 2005).
Geochemistry and Attenuation Processes
Radioactive Decay
According to documentation in Tuli (2005), there are 37
radioisotopes of radium, four of which occur naturally as a
part of the decay series for 238U, 235U, and 232Th. Following
are the four natural radioisotopes of radium, along with their
respective half-life and natural decay progenitor (see also
Uranium and Thorium chapters):
• 223Ra has a half life of 11.434 days and is part of the
235U decay series,
• 224Ra has a half life of 3.66 days and is part of the 232Th
decay series,
• 226Ra has a half life of 1600 years and is part of the
238U decay series, and
• 228Ra has a half life of 5.75 years and is part of the
232Th decay series.
Since 232Th is produced in the 236U decay series, the rela-
tive abundance of 224Ra and 228Ra may be increased, rela-
tive to typical natural abundance, for sites where enriched
uranium sources contribute to the ground-water plume.
Likewise, elevated levels of 223Ra may also be observed for
sites where 235U from uranium enrichment and processing
activities are a source for contaminant plume development.
These decay series are illustrated in Figure 8.1. Due to its
relatively longer half life and production within the 238U decay
series, 226Ra is typically the most abundant radioisotope in
natural systems. The decay of the four naturally-occurring
radioisotopes of radium produces radioisotopes of radon
as follows: 1) 228Ra decay produces 220Rn (or "thoron") via
224Ra (see Figure 8.1), 2) 226Ra decay produces 222Rn (or
"radon"; see Figure 8.2), and 3) 223Ra decay produces
219Rn (or "actinon"; see Figure 8.1). The energy of alpha
particles produced during thorium decay is sufficient to
eject the radium radioisotopes from the surface of solid
matrices (Cowart and Burnett, 1994). As discussed by
Vinson et al. (2009), alpha recoil is a likely physical mecha-
nism that enhances the mobility of radium. This process
will compete with chemical processes, such as adsorption
and coprecipitation, which serve to limit radium mobility
(discussed below).
79
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236JJ
2.34x1-). The half-life is shown directly
below the isotope that is subject to decay;
y = years, d = days, h = hours, m = minutes, s
= seconds. Numbers shown in parentheses
below indicate the fractional abundance of the
daughter isotope produced during branched
decay of the parent isotope (e.g., decay of227Ac
follows two routes resulting in 98.6% production
of227Th and0.014% production of223Fr). Decay
half-life data were obtained using the WinChain
program that provides electronic access to the
ICRP38 Nuclear Decay Data Files (ICRP, 1983;
Eckerman et ai, 1994; m = minutes, h = hours,
d=days, y=years). WinChain is a public domain
software application available for download from
Oak Ridge National Laboratory
(http://ordose. ornl. gov/dow nloads.html).
226Ra
1.60x103y
226Ra
Decay
Series
222Rn
3.82 d
a
i
1.5s
(0.0002)
#
218p0
3.05m
a
i
7
26.8m
(0.9998)
a
ir
"
214Bj
19.9m
X
P-
214Po
164 MS
1
P-
i
a
'
22.3 y
™
210Bj
5.01 d
*
P-
210Ro
138 d
a
a
, •
stable
F i g u re 8.2 Decay series for 228Ra based
on data from ICRP (1983). Decay modes include
those leading to ejection of an alpha particle
(a.) or a beta particle ($-). The half-life is shown
directly below the isotope that is subject to
decay; y = years, d = days, m = minutes,
s = seconds, i^s = microseconds. Numbers
shown in parentheses below indicate the frac-
tional abundance of the daughter isotope
produced during branched decay of the parent
isotope (e.g., decay of 218Po follows two routes
resulting in 98.98% production of 214Pb and
0.02% production of218At). Decay half-life data
were obtained using the WinChain program
that provides electronic access to the ICRP38
Nuclear Decay Data Files (ICRP, 1983;
Eckerman et ai, 1994; m = minutes, h = hours,
d = days, y = years). WinChain is a public
domain software application available for down-
load from Oak Ridge National Laboratory
(http://ordose. ornl. gov/dow nloads. html).
80
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Aqueous Speciation
Radium is an alkaline-earth element. These elements
follow a number of trends that help predict their chemical
behavior in the aqueous and solid phases in the subsurface
environment. All alkaline-earth elements have +2 oxidation
states and their hydrated ionic radii increase in the order
Ra2+ < Ba2+ < Sr2+ < Ca2+ < Mg2+ (Huheey, 1983). The ther-
modynamic data of aqueous and solid speciation of radium
was reviewed by Langmuirand Riese (1985). Ra2+ has little
tendency to form aqueous complexes and of the alkaline-
earth metals, Ra2+ shows the least tendency for complex
formation (Langmuir and Riese, 1985). Furthermore, it
may be assumed that Ra2+ does not hydrolyze (Baes and
Mesmer, 1976). Throughout the pH range of 3 to 10, the
uncomplexed ion Ra2+ is expected to be the dominant
aqueous species for dissolved radium (Pardue and Quo,
1998; Sturchio et al., 2001; USERA 2004). As illustrated
by the analyses presented in evaluations of radium aque-
ous chemistry by Paige et al. (1998) and Sturchio et al.
(2001), the existence of aqueous complexes of radium,
such as RaSO4(aq), would likely only be significant under
extreme conditions such as may be present in sulf uric acid
leach solutions within contaminant source areas. Transport
of radium downgradient into more dilute ground-water
conditions would result in the conversion to Ra2+ as the
predominant aqueous species.
Solubility
Most radium compounds, including Ra(NO3)2, RaCI2, and
Ra(IO3)2, are very soluble. Precipitates of radium that are
sparingly soluble include RaCO3 and RaSO4 (Langmuir
and Melchior, 1985; Baker and Toque, 2005). In ground
water containing moderate to high sulfate concentrations,
radium can precipitate as RaSO4 or coprecipitate with
barium as (Ba,Ra)SO4 (Langmuir and Melchior, 1985;
Pardue and Quo, 1998), and, to a lesser extent, radium
may coprecipitate with calcium in gypsum (Beddow et
al., 2006; Yoshida et al., 2009). Likewise, radium may be
coprecipitated with calcium during calcite precipitation in
ground-water systems with elevated alkalinity (Yoshida et
al., 2008). Coprecipitation is expected to be more common
because ground-water concentrations of radium are typi-
cally too low to support pure phase precipitation of RaSO4.
As a point of reference, stability fields for potential sulfate
or carbonate precipitates of calcium or barium that may
coprecipitate radium are shown in Figure 8.3. In these plots,
the shaded regions show where these precipitate phases
would form and remain stable. In general, carbonates would
dominate at pH>7 in ground-water systems with elevated
alkalinity. In addition, these diagrams illustrate that barium
sulfate (barite) is much less soluble than calcium sulfate
(gypsum). These projections appear to be consistent with
observations made for some ground-water and sediment
systems (Pardue and Quo, 1998; Martin and Akber, 1999;
Grundl and Cape, 2006).
a)
*•*/
-2
0*-3
w
n>
D)
_0 -4
-5
fi
Gypsum
Ca"=10"
_____
Ca2'
/
Calcite
Alstonite
25°C
b)
-2
O -3
(/)
n
01
o ^
Barite
Ba"=10
Ba'
Ba"=105m
Witherite
Alstonite
25°C
456
8
PH
10
11
12
Figure 8.3 pH - log aSO/- diagrams showing stability
fields for minerals in the Ba-Ca-Ra-SO4-CO2-H2O system.
Solubility data for Ra carbonate were taken from Langmuir
and Riese (1985). Gypsum=CaSO4-2H2O; Calcite=CaCO3;
Alstonite=BaCa(CO3)2; Barite=BaSO4; Witherite=BaCO3.
a) PCO2=10-3 bars and variable Ca2+ from 1Q-3mto m18 m.
b) PCO2=m3 bars and variable Ba2+ from W7 mto 10~5 m.
At the MCL for Ra (2.2e-13 m), the Ra carbonate and Ra
sulfate end-members are highly undersaturatedat PCO2= 103
bars and log a SO/- from -6 to -1; however, Ra could be
sequestered by Ba and Ca sulfates and carbonates.
81
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Adsorption
Recent overviews of radium adsorption processes in sub-
surface systems have been provided in USEPA (2004) and
IAEA (2006). In general, it is assumed that ion exchange
reactions would control radium adsorption based on pat-
terns of uptake onto clay minerals (Tachi et al., 2001) and
the apparent reversibility of the adsorption process (Ames
et al., 1983; Centeno et al., 2004). For systems in which
ion exchange reactions dominate radium partitioning to
aquifer solids, the presence of competing cations in ground
water will limit the extent of radium uptake (Berry et al.,
1994). However, patterns in ground-water chemistry for
some field studies (Herczeg et al., 1988), as well as direct
observations in radium distributions in ground-water and
associated aquifer solids (Gonneea et al., 2008), indicate
that adsorption to iron/manganese oxides can also play a
significant role in radium attenuation. In most instances,
it is reported that radium adsorption onto iron/manganese
oxides increases with pH, which is commonly observed
for cations that form inner-sphere surface complexes on
these solids. Based on use of manganese oxides in the
pre-concentration of radium from water samples, it is sug-
gested that these minerals exert some degree of selectivity
for radium adsorption (Moore and Reid, 1973; see also
discussion under Aqueous Measurements). At present,
there is insufficient information to conclude that radium will
show a preference for adsorption onto manganese oxides
over iron oxides.
Hidaka et al. (2007) have observed excess accumulation
of 226Ra in the clay mineral, illite, within sandstone forma-
tions near the Oklo uranium deposit. Comparison of the
relative distributions of stable and radioactive isotopes of
radium, barium, lead, and uranium for illite and other aqui-
fer minerals (calcite and quartz) indicated that radium was
selectively taken up by illite, potentially as a result of ion
exchange within interlayer sites of this clay mineral. The
close correspondence between elevated concentrations of
both radium and barium with illite suggests that the ultimate
fate of adsorbed radium may have been coprecipitation of
barite. However, no information was presented on either
the concentrations of sulfur or sulfate with illite from these
aquifer solids, which would be needed to provide evidence
for coprecipitation in barite as the long-term immobilization
process. Alternatively, both radium and barium may have
been selectively retained on ion exchange sites that resisted
desorption over time in this aquifer system.
Site Characterization
Attenuation of radium might be achieved through copre-
cipitation or adsorption dependent on the prevailing
ground-water chemistry within the plume and the relative
abundance and stability of immobile sorbent phases asso-
ciated with aquifer solids. While radioactive decay may
be a viable process for the short-lived radioisotopes, i.e.,
223Ra 224Ra and 228Rg thjg wj|| |jke|y not gerve ag a vjab|e
attenuation process for plumes dominated by the longer-
lived 226Ra. Co-precipitation reactions are likely to occur
only under situations where elevated sulfate concentrations
drive precipitation of minerals such as barite or gypsum.
Adsorption onto iron/manganese oxides and clay minerals
will likely be the dominant attenuation process for aquifers
undersaturated with respect to precipitation of sulfate/car-
bonate minerals. A list of potential attenuation processes
is provided in Table 8.1. Two factors that will dictate the
adequacy of attenuation via coprecipitation or adsorption
include the rate of water transport and the total mass and
release rate of radium into the subsurface plume. Details
of the types of analytical measurements that may be con-
ducted on sampled ground water and aquifer sediments
to assist in identifying the attenuation mechanism(s) are
discussed in the following paragraphs.
Table 8.1 Natural attenuation and mobilization pathways for radium.
Attenuation Processes
Mobilization Processes
Characterization Approach
Coprecipitation with sulfate
minerals such as barite or
gypsum
Dissolution due to decreasing ground-
water sulfate concentrations and/or
microbially driven sulfate reduction
Evaluate formation of sulfate
minerals along relevant transport
pathways and determine radium
association with this mineral fraction.
Adsorption or ion exchange
onto aquifer minerals (clay
minerals, iron/manganese
oxy hydroxides)
Desorption due to decreasing pH or
competition from major cations in
ground water; reductive dissolution of
iron/manganese oxyhydroxides.
Evaluate total adsorption capacity of
aquifer solids under representative
ground-water chemistry; chemical
extractions to assess concentrations
of exchangeable radium fractions
in aquifer solids along relevant
transport pathways.
82
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Aqueous Measurements
Due to its longer half-life, 226Ra is the primary radioisotope
encountered in ground-water plumes developed from the
disposal of radium-bearing wastes with distributions of
uranium and thorium parent radioisotopes similar to those
found in natural systems. However, analysis of 228Ra,
223Ra, and/or 224Ra may be warranted for situations in
which delineation of the source(s) of ground-water radium
contamination is needed (e.g., contaminant source zones
that may have elevated levels of 235U or 236U parents; Martin
and Akber, 1999). Measurement of the distribution of radium
radioisotopes has been used as a marker for radium-
enriched ground-water discharge zones into surface water
(e.g., Lauria and Godoy, 2002). As shown above, decay of
radium radioisotopes involves emission of an alpha particle
(predominant energies: 226Ra, 4.60 and 4.78 MeV; 223Ra,
5.54, 5.61, 5.72, and 5.75 MeV; 224Ra, 5.45 and 5.69 MeV)
or beta particle emission (228Ra, 12-40 keV), and, thus, can
be measured using radiometric techniques (USEPA, 2006;
Hou and Roos, 2008). Analysis of radium in environmental
samples generally is targeted towards detection of 226Ra
and 228Ra due to relative abundance. For both of these
radioisotopes, it is common to determine activity indirectly
via quantification of their first progeny, 222Rn for 226Ra and
228Ac for 228Ra (USEPA, 2006). This approach typically
entails chemical separation of the parent/daughter radio-
isotope from the sample matrix to avoid interference from
other radioisotopes that may be present (e.g., 230Th, 234U,
90Sr) followed by a period of daughter in-growth to establish
secular equilibrium between the parent and its daughter
(USEPA, 2006; Hou and Roos, 2008). Due to similarity in
chemical properties, 133Ba is typically employed as a tracer
for determining chemical recovery during separation of
radium isotopes from the sample matrix. In general, these
radiometric techniques provide typical detection limits of 1
pCi/L for both 226Ra (1 x10'9 mg/L) and 228Ra (3.7x10~12 mg/L)
(USEPA, 2006). If analysis of the short-lived radioisotopes
223Ra and 224Ra is planned, then the isolation and quan-
tification should be conducted as soon as possible after
sample collection. Otherwise, loss of parent and daughter
radioisotopes due to decay during sample holding may
result in activities below analytical detection.
Radium radioisotopes may be separated from the ground-
water matrix via chemical extraction, coprecipitation with
lead/barium sulfate (Bandong et al., 2005) or manganese
dioxide, or through use of ion exchange or chelating resins
(Rihs and Condomines, 2002; Lariviere et al., 2005; Aguado
et al., 2008). Chemical extraction is commonly employed
in combination with detection using liquid scintillation due
to the availability of liquid scintillants that are efficient for
radium extraction (e.g., Aupiais, 2005). Separation via
coprecipitation with or adsorption onto manganese dioxide
is frequently used to concentrate radium from water sam-
ples prior to a wide range of detection methods (Eikenberg
et al., 2001; Dulaiova and Burnett, 2004; Ghaleb et al., 2004;
Nour et al., 2004; Zoriy et al., 2005; Karamanis et al., 2006;
Peterson et al., 2009), and there have been detailed studies
to evaluate the various analytical parameters that influence
the efficiency of radium extraction from water samples using
this separation method (Eikenberg et al., 2001; Dimova
et al., 2008; Garcia-Solsona et al., 2008; Moore, 2008).
Pre-concentration of radium onto manganese dioxide has
also been employed in field measurement methods that
provide indirect quantification of radium radioisotopes based
on detection of progeny such as radon (Kim et al., 2001;
Dimova et al., 2007).
For the longer-lived 226Ra, mass-based detection tech-
niques, using ICP-MS as the analytical platform, have
become increasingly reliable and sensitive. Hou and Roos
(2008) have reviewed issues of potential analytical interfer-
ence from the sample matrix, many of which can be largely
controlled with the advent of improved sample introduction
systems and collision-cell technologies for the interface
between the plasma and the mass spectrometer (Lariviere
et al., 2006). As with radiometric techniques, separation
of 226Ra from the matrix improves analytical detection and
interference from the sample matrix (Lariviere et al., 2003).
Recent developments have made it possible to incorporate
and automate the separation procedure in-line with the
ICP-MS system (Lariviere et al., 2003; Benkhedda et al.,
2005), thus achieving improvements in sample throughput.
In general, the sensitivity of mass-based detection is cur-
rently insufficient for this technique to compete with the
radiometric techniques described above for ground water
where detection of low concentrations may be needed.
[It should be noted that use of drilling fluids
containing barite should be avoided during ground-
water well installation due to the potential bias
resulting from dissolution of naturally-occurring
radium associated with barite in residual drilling
fluids that can be difficult to remove from the
screened interval. Due to the half-lives of the
various radium radioisotopes, 22eRa would be the
most likely source of contamination from barite in
drilling fluid.]
Solid Phase Measurements
Solid phase measurements that may provide information
useful to assessing processes controlling radium retardation
within the aquifer and the capacity along relevant trans-
port pathways include the determination of partitioning of
radium radioisotopes to aquifer solids, the cation-exchange
capacity of aquifer solids, and identification of aquifer sol-
ids mineralogy that may participate in adsorption and/or
coprecipitation reactions. Evaluation of the mass distribu-
tion of radium radioisotopes between co-located ground
water and aquifer solids throughout the plume provides
an assessment of the extent that retardation reactions
limit radium migration. Different approaches have been
applied to the determination of the solid-phase speciation
of radium in aquifer solids. One approach involves looking
for correlations between radium and other elements within
aquifer solids as a function of grain size or density frac-
tions using either non-destructive (i.e., X-ray fluorescence
for stable elements and alpha/gamma spectrometry for
83
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radioisotopes of radium) or destructive techniques based
on decomposition of the solid sample with strong acids or
alkaline fusion. Iwaoka et al. (2009) describe approaches
for analyzing ore materials using acid or alkaline fusion
decomposition methods prior to mass-based determination
of stable element concentrations and gamma spectrom-
etry to determine total 226Ra and/or 228Ra concentrations
via detection of 214Bi and 228Ac progeny, respectively. For
decomposition methods that solubilize radium into a liquid
matrix, methods discussed under the section on Aqueous
Measurements tor sample preparation and analysis may be
applicable. Ghaleb et al. (2004) and Lariviere et al. (2007)
discuss approaches to pre-condition extract samples prior
to analysis in order to remove high concentrations of matrix
elements and/or radionuclides that might interfere with
analytical quantification of radium. As described by Saint-
Fort et al. (2007), it may be possible to use field-screening
detection methods, based on measurements of total radia-
tion with field meters, for the purpose of targeting samples
for more detailed laboratory characterization.
Since radium attenuation will likely involve partitioning to
existing or newly formed solid phases within the aquifer,
chemical extraction procedures are commonly employed
to target specific phases. For systems in which ion-
exchange onto clay minerals is suspected, aquifer solids
may be extracted using a concentrated salt solution with
a cation that would displace radium from the exchange
sites. An assumption of these procedures is that only
exchangeable cations are released from the solid matrix.
It is recommended that the potential for partial dissolution
of the more labile mineral fraction in aquifer solids during
extraction be assessed (e.g., Jackson and Inch, 1989).
This may be achieved through measurement of the major
ion chemistry in the extract solution (e.g., iron, manganese,
sulfate), which may serve as markers for the dissolution
of oxyhydroxides, sulfates or other components that may
bind radium in a less labile form. Solutions containing a
strong chemical reducing agent are commonly employed to
target iron/manganese oxides to which radium (or barium)
may be adsorbed or coprecipitated (Aguado et al., 2004;
Charette et al., 2005). Benes et al. (1981) and Rutten et
al. (2002) describe procedures for selective dissolution of
barite using solutions of EDTA in ammonium hydroxide or
an ammonium chloride solution at pH 7, respectively. For
aquifer solids in which it is feasible to selectively dissolve
radium-bearing barite, it may be possible to measure the
ratio of extracted 228Ra and 226Ra relative to the concentra-
tion of these radioisotopes in co-located ground water, to
determine whether solid-phase radium is a background
component of the aquifer solids (i.e., 226Ra predominates)
or a newly-formed component (i.e., 228Ra and 226Ra reflect
ground-water composition) derived from attenuation of
radium within the ground-water plume (e.g., Zielinski et al.,
2001). Alternatively, radium association with barite or other
relatively insoluble phases may be assessed by selective
extraction of other potential sorbent mineral phases prior
to examination of radium associated with residual solids.
Several studies have shown that radium coprecipitated
with barite or strongly bound to clay minerals may not be
extracted by solutions designed to target radium parti-
tioned to readily accessible ion exchange sites, weak acid
extractable phases (e.g., calcite), and/or reducible iron/
manganese oxides (Pardue and Quo, 1998; Hidaka et al.,
2007; Leopold et al., 2007).
Determination of the host mineral phase(es) dissolved for
each extraction step is recommended, along with the use
of surrogate radium-bearing phases spiked into the sedi-
ment to confirm accuracy of the procedure (e.g., Rudd et
al., 1988). The choice of appropriate radium surrogate
phase(s) would be governed by site-specific geochemi-
cal conditions or characterization of the mineralogy of the
aquifer sediment. Based on information for likely endpoints
for solid-phase partitioning, radium-bearing surrogate
phases may include radium-bearing barite (natural or
synthesized), radium-bearing gypsum or calcite, radium
adsorbed to or coprecipitated with iron/manganese oxides,
or radium-exchanged clay minerals. Spiking aquifer solids
with these surrogate phases provides the basis for assess-
ing the actual selectivity of the various extraction solutions
used to target specific solid components. Physical and
chemical procedures, e.g., size fractionation, may also be
employed to assist in identifying specific components within
aquifer solids that dominate radium solid-phase partitioning.
Examples of procedures to identify the type and abundance
of specific minerals along transport pathways are available
in the literature (e.g., Amonette, 2002). Additional infor-
mation on analysis approaches and analytical techniques
applied to solid phase characterization is provided in USEPA
(2007; Section NIB).
Long-Term Stability and Capacity
The long-term stability of radium attenuated through copre-
cipitation or adsorption will depend upon the stability of the
host mineral and the abundance of other ions which might
displace adsorbed radium. The most easily envisioned case
of the first is that radium coprecipitated in sulfate minerals
might be remobilized if sulfate-reducing conditions develop
(e.g., Pardue and Quo, 1998; Landa, 2003; Martin et al.,
2003) due to a decreased influx of oxygen as terminal
electron acceptor and/or an increased influx of readily
degradable organic compounds in excess of the capacity of
otherterminal electron acceptors [e.g., solid phase Fe(lll) or
Mn(IV)] consumed as a result of stimulated microbial activ-
ity. Dissolution of host sulfate precipitates may also result
from a decrease in ground-water sulfate concentration (e.g.,
Pulhani et al., 2007). In addition, reductive dissolution of
manganese and/or iron oxyhydroxides might cause remo-
bilization of adsorbed radium (e.g., Herczeg et al., 1988;
Landa, 2003). Increases in divalent cation levels can also
be expected to work against radium immobilization for sites
dominated by cation exchange reactions (e.g., Martin and
Akber, 1999; Sturchio et al., 2001). Review of the extent
of plume development for a number of sites indicates that
cation exchange within the saturated aquifer may have
insufficient stability to prevent plume expansion (Brady et
al., 2002). Thus, a critical factor for assessing the overall
84
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capacity of the aquifer for attenuation will be evaluation of
the mass flux of dissolved radium moving through the plume
relative to the rate of water movement through the aquifer.
Translocation of 226Ra through the aquifer due to changes in
chemistry may also influence the dimensions and location
of a co-occurring 222Rn plume. This may pose a concern at
sites with shallow ground-water systems where exposure
to radon in indoor air and/or exposure of ecological/human
receptors in surface water following ground-water discharge
may present potential contaminant exposure pathways. In
order to make a reliable assessment of the mass/activity
flux of radon throughout the plume, it will be important to
understand the characteristics of the hydrogeologic system
and the dynamics of water and radium transfer throughout
the plume (see Radon chapter).
Tiered Analysis
Determination of the viability of radium remediation in
ground water via monitored natural attenuation will depend
upon proper assessment of contaminant loading to the
aquifer relative to the velocity of ground-water travel and
the prevailing geochemistry and mineralogy within the
contaminant plume and the down gradient zone prior to the
point(s) of compliance. The goal of site assessment is to
demonstrate the process(es) controlling 226Ra sequestra-
tion onto aquifer solids and the long-term stability of solid
phase radium as a function of existing and anticipated
ground-water chemistry. If short-lived 223Ra, 224Ra, and/
or 228Ra are a component of the plume, then radioactive
decay may serve as a primary attenuation mechanism for
these radio isotopes. The following tiered analysis structure
for site characterization provides an approach to evaluate
candidate sites and define the potential limitations of MNA
as part of a remedy for ground-water cleanup.
Tier I. Site characterization under Tier I will involve demon-
stration that the ground-water plume is static or shrinking,
has not reached compliance boundaries, and does not
impact existing water supplies. Once this is established
through ground-water characterization, evidence is col-
lected to demonstrate radium partitioning to aquifer solids
within the plume. If immobilization processes are active
throughout the plume, then there should be an observed
increase in solid phase concentrations within regions of the
plume with higher aqueous concentrations, e.g., near the
source term. Evaluation of the mass/activity of radium dis-
tributed between ground water and aquifer solids throughout
the plume is recommended to account for both existing
and potentially mobile forms of radium. This field partition-
ing data may be supplemented by geochemical modeling
that incorporates measured water chemistry (e.g., pH and
concentrations of barium, calcium, and sulfate) throughout
the plume to assess the potential for solubility control by
coprecipitation with sulfate minerals. It is also important
at this stage of the site evaluation to determine source
term characteristics such as the inventory of contaminant
mass and the current and historical rate of release into the
subsurface. Acquisition of this information in combination
with identification of a stable plume provides justification
for proceeding to Tier II characterization efforts.
Tier II. Under Tier II, the apparent rate and mechanism(s)
of attenuation are determined. Estimates of a site attenu-
ation rate(s) can be assessed via a well transect along
the ground-water flow path. In addition, time-series data
may be collected at one or more monitoring points within
the plume (USEPA, 2007; Section IIIA.5). This information
will allow assessment of the relative timescales for con-
taminant attenuation and fluid transport and determination
of whether remediation objectives can be met within the
required regulatory time frame. As part of this analysis, it
is critical to determine the relative contributions of radio-
active decay (for short-lived 223Ra, 224Ra, and/or 228Ra)
and immobilization to the overall observed attenuation.
Determination of the contribution of radioactive decay will
necessitate detailed analysis of system hydrology relative
to flow pathway(s), flow velocity, and temporal variations in
flow velocity and/or direction within the boundaries of the
plume. This information, in combination with knowledge
of contaminant source release characteristics, can be
employed to develop a decay-transport model to project
radionuclide activity/concentration distribution throughout
the plume in the absence of adsorption/coprecipitation
processes. For systems in which immobilization plays a
role in observed attenuation, it will be necessary to identify
whether adsorption onto existing aquifer minerals or copre-
cipitation with newly formed minerals predominates. This
effort will require determination of the chemical speciation
of solid phase 226Ra and may be approached according to
the following scheme:
1. Calculation of saturation state of ground water rela-
tive to precipitation of sulfate minerals along relevant
radium transport pathways;
2. Determination of aquifer mineralogy to determine the
relative abundance of components with documented
capacity for radium adsorption with implementation
of steps for aquifer solids collection, processing and
analysis that avoid transformation of mineral species
from reduced zones (e.g., oxidation of ferrous carbon-
ate to ferric (hydr)oxide); and
3. Determination of solid phase radium associations
via chemical extractions designed to target specific
components within the aquifer solids.
This compilation of information will facilitate identifica-
tion of the reaction(s) leading to radium attenuation. The
demonstration of concurrence between conceptual and
mathematical models describing radium transport will entail
development of site-specific parameterization of the chemi-
cal processes controlling radium solid phase partitioning.
Tier III. Once the contributions from radioactive decay
and adsorption/coprecipitation processes have been deter-
mined, the subsequent characterization effort under Tier
III will involve determination of the stability of immobilized
radium and the capacity of the aquifer to sustain continued
uptake. It is recommended that the stability of immobilized
radium be tested based on the anticipated evolution of
ground-water chemistry concurrent with plume evolution.
For example, changes in ground-water pH and/or cation
composition can exert a significant influence on radium
85
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adsorption. Therefore, it is recommended that sediment
leach tests be conducted to characterize the magnitude
of radium re-mobilization as a function of pH for a ground-
water chemistry representative of site conditions. It is rec-
ommended that the capacity for radium uptake onto aquifer
solids be determined relative to the specific mechanism(s)
identified in Tier II. For sites in which a continuing source
of radium to the saturated aquifer exists, it is recommended
that potential steps to minimize or eliminate this continued
contaminant flux be evaluated and implemented where
feasible. If site-specific tests demonstrate that the stabil-
ity and capacity for radium immobilization are sufficient to
sustain attenuation, then the site characterization effort can
progress to Tier IV.
Tier IV. Finally, under Tier IV a monitoring plan is estab-
lished along with contingency plans in the event of MNA
failure. It is recommended that the monitoring plan be
designed to establish both continued plume stability and
to identify changes in ground-water chemistry that may
lead to re-mobilization of attenuated radium. The specific
chemical parameters to be monitored will include those
identified under Tier III that may halt radium partitioning to
aquifer minerals and/or result in solubilization of precipitates
into which radium has been incorporated. Solution phase
parameters that could alter either radium coprecipitation or
adsorption include changes in pH and sulfate concentra-
tions and/or increased concentrations of competing cations
in ground water. As an example, increases in barium con-
centrations in ground water could signal either 1) the poten-
tial for displacement of radium from cation exchange sites,
or 2) the dissolution of barium sulfate minerals in which
radium is coprecipitated. Likewise, increases in Fe(ll),
Mn(ll), and/or dissolved sulfide concentrations may signal
the onset of reducing conditions that are dissolving aquifer
solid phases to which radium is partitioned (e.g., sulfate
precipitates or iron/manganese oxides). Changes in water
chemistry may occur prior to observed changes in solution
radium activity/concentration and, thus, serve as monitor-
ing triggers for potential MNA failure. In this instance, a
contingency plan can be implemented that incorporates
engineered strategies to arrest possible plume expansion
beyond compliance boundaries. Possible strategies to
prevent plume expansion include ground-water extraction
with surface treatment, installation of permeable reactive
barriers to enhance uptake capacity perpendicular to the
direction of plume advance, or enhancement of coprecipi-
tation processes within the aquifer through the injection of
soluble reactive components.
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Thorium
Daniel I. Kaplan, Robert G. Ford, Richard T. Wilkin
Occurrence and Distribution
Thorium rock content varies greatly depending on the rock's
origin (Adam and Richardson, 1960; Deer et al., 1967; Ames
and Rai, 1978). Thorium concentrations in sedimentary
rocks range from 1.1 mg/kg in limestone to 49 mg/kg in
bauxite. It concentrates in sedimentary rocks primarily by
selective adsorption on clays and coprecipitation in heavy
resistate minerals, such as monazite [(Ce,La,Th)POJ or
zircon (ZrSiO4). Concentrations of thorium in metamorphic
rocks are especially variable, ranging from 0.03 mg/kg in
marble to 13.1 mg/kg in gneiss. Thorium concentrations
in igneous rocks increase along the petrologic series from
basalts (0.5 to 2 mg/kg) to granites (10 to 20 mg/kg). Finally,
thorium concentrations in unimpacted sediments vary from
3.8 to 12.4 mg/kg (Govindaraju, 1994). Most thorium-
containing minerals are sparingly soluble and resistant to
erosion (Frondel, 1958). Thorium is an essential structural
constituent in few minerals. Instead thorium usually occurs
in variable concentrations in solid solution in rare-earth,
zirconium, and uranium minerals. Thorium exists in com-
mercial concentrations (percent levels) in detrital sands
containing monazite and zircon (e.g., Baxter, 1996; Powell
et al., 2007). Since long-lived thorium radioisotopes derive
from decay of naturally-occurring radioisotopes of uranium,
elevated subsurface concentrations of thorium may be
encountered at sites where uranium mining, uranium ore
milling, or uranium nuclear fuel production and nuclear
fuel reprocessing has taken place (Carvalho and Oliveira,
2007; USEPA, 2008).
Thorium content in natural streams is very low, rarely
exceeding 1 |ag/L or 0.1 pCi/L 232Th (Hem, 1985), and it
is even lower in oceans, averaging ~5.3 x 10'5ug/L (Huh
and Bacon, 1985). As will be discussed below, these con-
centrations are low because thorium partitions strongly to
essentially all mineral surfaces and it readily forms sparingly
soluble precipitates, such as thorianite (or hydrous thorium
oxide), ThO2, in many ground-water systems (Langmuir
and Herman, 1980). However, some naturally high thorium
concentrations in ground water exist. Thorium concen-
trations in excess of 2000 |ag/L were reported in acidic
ground waters beneath uranium tailings sites (Langmuir
and Herman, 1980). In another example, natural aqueous
thorium concentrations of >22 |ag/L were detected in Soap
Lake in eastern Washington (LaFlamme and Murray, 1987).
Geochemistry and Attenuation Processes
Radioactive Decay
There are 26 radioisotopes of thorium, 14 of which
have half-lives less than one second. Thorium iso-
topes that may be of concern for waste disposal opera-
tions include 227Th (t1/2 = 18.718 days), 228Th (t1/2 =
1.9131 years), 229Th (t1/2 =7340 years; Figure 9.1), 230Th
(t1/2= 77,000 years; Figure 9.2), 231Th (t1/2= 25.52 hours),
232Th (t1/2= 1.405 x 1010 years), and 234Th (t1/2 = 24.1 days).
The 229Th radioisotope is produced from decay of man-made
233U, which is proposed for use to produce the 213Bi radio-
isotope for medical therapy applications (e.g., Apostolidis et
al., 2005). However, while there is a limited inventory of 233U
(t1/2 = 1.585 x 105 years) at some DOE sites, the amount of
229Th is not considered to be a significant issue for ground
water. Two thorium radioisotopes, 234Th and 230Th, occur
naturally as decay products within the 238U decay series.
The 231Th and 227Th radioisotopes are produced within the
235U decay series, and 228Th is a decay product within the
232Th decay series (Figure 9.3). Of the three longest-lived
thorium radioisotopes, 232Th is the most abundant in nature
(>99%), and it may also be derived from decay of 236U (t1/2
= 2.341 x 107 years) at sites where spent nuclear fuel was
disposed or reprocessed. As can be seen in Figures 9.1-
9.3, radioisotopes of radium and radon are produced in the
decay series for 229Th, 230Th, and 232Th.
91
-------
2381J
4.47x10'y
a
234pg
6.70 h
234U
2.44x10*y
a
230Th
7.70x1 0*y
oc
1.60x1 0s y
a
a
218Ro
3.050 m
a
214pb
26.8m
234Th & 230Th
Decay
Series
214Bi
19.9m
164ns
210pb
22.30 y
210B|
5.01 d
210p0
138d
a
6pb
Figure 9.1 Decay series for229Th including parent233 U. The
half-life is shown directly below the isotope that
issubjecttodecay;y=years, d=days, h=hours,
m = minutes, s = seconds, \ns = microseconds,
stable = non-radioactive isotope. Decay half-
life data were obtained using the WinChain
program that provides electronic access to
the ICRP38 Nuclear Decay Data Files (ICRP,
1983; Eckerman et ai, 1994). WinChain is a
public domain software application available for
download from Oak Ridge National Laboratory
(http://ordose. ornl. gov/downloads. html).
Figure 9.2 Decay series for 234Th and 230Th including
parents 238U and234U. Production of226Ra and
"radon" (222Rn) within this decay series is high-
lighted. The half-life is shown directly below the
isotope that is subject to decay; y = years, d =
days, h = hours, m = minutes, |as = microsec-
onds, stable = non-radioactive isotope. Decay
half-life data were obtained using the WinChain
program that provides electronic access to
the ICRP38 Nuclear Decay Data Files (ICRP,
1983; Eckerman et ai, 1994). WinChain is a
public domain software application available for
download from Oak Ridge National Laboratory
(http://ordose. ornl. gov/downloads.html).
92
-------
232Th & 228Th
Decay
Series
Figure 9.3 Decay series for 232Th including parent 236U
andprogeny228Th. Production of228Ra, 224Ra,
and "thoron" (220Rn) within this decay series
is highlighted. The half-life is shown directly
below the isotope that is subject to decay; y
= years, d = days, h = hours, m = minutes,
s = seconds, [is = microseconds, stable =
non-radioactive isotope. Decay half-life data
were obtained using the WinChain program
that provides electronic access to the ICRP38
Nuclear Decay Data Files (ICRP, 1983;
Eckerman et al., 1994). WinChain is a pub-
lic domain software application available for
download from Oak Ridge National Laboratory
(http://ordose. ornl. gov/downloads.html).
Aqueous Speciation
Although other oxidation states of thorium can exist under
extreme laboratory settings, onlyTh(IV) is found in nature.
The atomic radius of Th+4 is 9.9 nm, and because of its
very small size and high charge it has an exceptionally
high effective charge that undergoes extensive interaction
with water and many anions. The available thermodynamic
data for thorium aqueous species and solids have been
compiled and critically reviewed (Hala and Miyamoto, 2007;
Rand et al., 2008).Thorium exists as a complex, i.e., a spe-
cies other than Th4+, in solutions of pH >3.5 (Figure 9.4a).
Above pH 3.5, thorium readily undergoes hydrolysis to
form a number of species, of which Th(OH)|+, Th(OH)+,
and Th(OH)°(ac7) are the most predominant (Neck and
Kim, 2001). Polynuclear complexes, in which more than
one Th(IV) atom is incorporated within the structure of the
aqueous complex, may also form at higher thorium con-
centrations (Rothe et al., 2002; Tsushima, 2008; Walther
et al., 2009). It is this tendency to form polymeric species
that is cited as a potential form of colloidal species that
may be transported (see Solubility).
A ranking of inorganic anions in order of their increas-
ing tendency to form complexes with thorium are:
nitrite < chloride < phosphate < sulfate < fluoride (Langmuir
and Herman, 1980). There is also evidence that the
presence of synthetic chelating agents, such as citrate,
EDTA or NTA in liquid waste streams, may also influence
the speciation of thorium near contaminant source zones
(Felmy et al., 2006; Cartwright et al., 2007). Under rela-
tively acidic conditions in ground water with high sulfate
concentrations, Th(IV) complexation with sulfate may
dominate its aqueous speciation (Figure 9.4b; Felmy and
Rai, 1992; Rand et al., 2008). Historical observations of
elevated thorium concentrations in waters with elevated
alkalinity have been attributed, in part, to the formation
of aqueous complexes with carbonate (e.g., Anderson
et al., 1982; LaFamme and Murray, 1987). Aqueous car-
bonate has been reported to form strong complexes with
thorium to form ternary species with the general species
formula Th(OH) (CO3)4-y-2z (Osthols et al., 1994; Altmaier
et al., 2005; Altmaier et al., 2006; Rand et al., 2008). The
potential influence of dissolved carbonate on the aqueous
speciation of Th(IV) is illustrated in Figure 9.5 for a system
with elevated sulfate. For these equilibrium calculations,
the partial pressure of CO2 was varied from 0.0001 to 0.01
atmosphere, which results in an increase in the stability
field of the ternary aqueous species, Th(OH)2(CO3)2-, and
a concomitant shrinkage in the stability fields for both solid
phase ThO2(am) and aqueous Th(OH)4. It is anticipated
that the formation of aqueous carbonate complexes may
decrease thorium adsorption onto aquifer minerals for
systems with elevated alkalinity (see Solubility).
Based on laboratory studies, complexation of thorium by
natural forms of dissolved organic carbon (DOC) such as
humic/fulvic acids may be expected to occur (Nash and
Choppin, 1980; Reiller et al., 2008; Benes, 2009). In gen-
eral, it has been proposed that formation of complexes with
dissolved natural organic matter could dominate thorium
93
-------
speciation in ground water with elevated concentrations of
DOC (e.g., Langmuir and Herman, 1980), although there
are very few published data for subsurface systems to
support this notion. Gaffney et al. (1992) provided direct
analysis data for pore water sampled from a shallow wet-
land system that demonstrate thorium association with
DOC for filtered samples. Also, Dia et al. (2000) present
ground-water data for an aquifer flow path along a hill slope
terminating in a wetland-riverine area that support a positive
relationship between DOC and thorium (filtered, <0.2 |am).
Characterization of the size distribution of DOC and tho-
rium concentrations following further size-fractionation of
<0.2 jam-filtered ground water below a wetland system
suggests a potential strong association between thorium
and colloidal forms of natural organic matter (Pourret et
al., 2007). However, as demonstrated by Hassellov et al.
(2007), conclusions about the apparent colloidal nature of
particle-reactive contaminants such as thorium should be
viewed with caution when strict controls are not taken at the
point of sampling to prevent intrusion of oxygen in suboxic
or anoxic ground-water samples. For example, precipitation
of ferrous iron at concentrations as low as 3 mg/L poten-
tially confound these interpretations, since both thorium
and DOC will tend to partition to these colloidal solids that
form during sample holding of un-acidified samples prior to
commonly employed fractionation procedures. Thus, while
existing chemical data support the potential for thorium to
migrate with dissolved or colloidal forms of DOC, there is
inadequate field data to support that this process can result
in facilitated transport of thorium over large distances in
an aquifer.
Solubility
Dissolution, precipitation, and coprecipitation are important
processes controlling aqueous thorium concentrations.
Current reviews of thermodynamic data describing the
solubility of pure precipitates of thorium are available in
Hala and Miyamoto (2007) and Rand et al. (2008). The
main thorium-containing minerals are resistant to chemical
weathering and do not dissolve readily at low temperatures
in surface and ground waters. Hydrous thorium oxide, ThO2,
has been shown to precipitate in laboratory experiments
conducted at low temperatures (Ryan and Rai, 1987). If
this solid forms in nature, it is likely that it would become
increasingly more crystalline and increasingly less soluble
(Fanghanel and Neck, 2002). As noted above, thorium
tends to form polymeric hydrolysis species, and this process
has been observed to cause elevated apparent soluble
thorium concentrations in pure aqueous suspensions of
hydrous thorium oxide (Fanghanel and Neck, 2002). It has
been proposed that these colloidal thorium species could
influence concentrations in ground-water, although there
are no field observations to verify the importance of this
process. Osthols et al. (1994) determined that the solubil-
ity of ThO2 increases greatly in the presence of dissolved
carbonate. As can be seen in Figure 9.5, the range of pH
at which aqueous concentrations of Th(IV) are controlled
by ThO2 solubility shrinks significantly with increasing dis-
solved carbonate concentrations (i.e., 10~4 < PCO2 < 10~2).
Adsorption
The partitioning of thorium to a range of soils and aquifer
solids has been presented in USEPA (1999). In general,
this review supports the observation of the tendency for
thorium to partition to aquifer solids. Under near-neutral
pH conditions, thorium will primarily exist as a neutral dis-
solved species (see Figure 9.4) and has been observed
to adsorb onto oxide and silicate minerals over a wide pH
range with significant adsorption observed even under
acidic conditions (e.g., Reiller et al., 2002; Bradbury and
Baeyens, 2009). Several studies have demonstrated that
thorium adsorption to mineral surfaces involves the for-
mation of chemical bonds with surface functional groups
(Dahn et al., 2002; Seco et al., 2009), and these reactions
can be described using surface complexation models (e.g.,
Degueldre and Kline, 2007; Bradbury and Baeyens, 2009).
Laboratory studies examining thorium partitioning to a wide
range of mineral components reveal that thorium adsorption
occurs on all types that might be encountered in aquifers
(e.g., Quo et al., 2002; Geibert and Usbeck, 2004; Santschi
et al., 2006). The adsorption of thorium to mineral surfaces
may be inhibited (or desorption increased) with increasing
concentrations of carbonate or dissolved organic matter in
ground water (e.g., Laflamme and Murray, 1987; Reiller et
al., 2002). As shown by Reiller et al. (2002), the influence
of dissolved organic matter such as humic/f ulvic compounds
will likely depend on the relative concentration of dissolved
organic matter and available adsorption sites on aquifer
solids, with adsorption inhibition (or desorption) becom-
ing a factor when dissolved organic matter concentrations
exceed adsorption site concentrations. These authors also
demonstrated that thorium adsorption may increase under
conditions where humic/fulvic compounds partition to the
mineral surface. Thus, it is anticipated that thorium will
also partition to immobile forms of natural organic matter
that may be present in aquifer solids for shallow ground-
water systems.
94
-------
-5
a)
-6
CD
O)
O _
-10
Th02(am)
TrT
Th(OH)a(aq)
25°C
0 46
PH
10 12 14
-5
b)
05
O)
O _8
-9
-10
Th(SO4)
2 fa
ThO2(am)
Th(OH)«(aq)
25°C
10
12
14
PH
Figure 9.4 Solubility of ThO2(am) in (a) pure water and
(b) 1 mM SO/'. Diagrams constructed at
25°C using the LLNL thermodynamic data-
base (therm.com.v8.R6+) modified using
data reported in Altmaier et al. (2006) for the
solubility of ThO2(am) and Rand et al. (2008)
for the solubility of Th(SO^-9HzO.
w
-6
r
^ -7
CD
O)
O -8
-9
-in
Th(SO4)2(a
° sTT
r r Is
0" o"' |»
P 0 l°
a a /<->
? r
s ^
° n
I
' /
/
L
"CT
-5-
f
O
(-
d'
o_
0
H
^_^
C3
O
0
I—
25°C
14
PH
Figure 9.5. Solubility of ThO2(am) over variable log PCO2
values (-4 to -2) and 1 mM SO/-. Diagram
constructed at 25°C using the LLNL thermody-
namic database (therm.com. v8. R6+) modified
using data reported in Altmaieretal. (2006) for
Th-OH-CO3 complexes and ThO2 (am) solubil-
ity, and Rand et al. (2008) for the solubility of
Th(SO4 )2-9H2O. The light gray lines show the
increase of the stability field for the aqueous
species, Th(OH)2(CO3)2- with increasing PCO2
Site Characterization
Attenuation of radium might be achieved through copre-
cipitation or adsorption dependent on the prevailing
ground-water chemistry within the plume and the relative
abundance and stability of immobile sorbent phases associ-
ated with aquifer solids. While radioactive decay may be a
viable process forthe short-lived radioisotopes, i.e., 228Th or
234Th, this will likely not serve as a viable attenuation pro-
cess for plumes dominated by the longer-lived 232Th. Due
to the low solubility of hydrous thorium oxide, precipitation
of this phase may occur near contaminant source areas
where elevated dissolved thorium concentrations may exist.
Adsorption onto clay minerals, iron/manganese oxides, and/
or immobile forms of natural organic matter will likely be the
dominant attenuation process for aquifers undersaturated
with respect to precipitation of thorium. A list of potential
attenuation processes is provided in Table 9.1. Two factors
that will dictate the adequacy of attenuation via precipitation
or adsorption include the rate of water transport and the
total mass and release rate of thorium into the subsurface
plume. Details of the types of analytical measurements
that may be conducted on sampled ground water and
aquifer sediments to assist in identifying the attenuation
mechanism(s) are discussed in the following paragraphs.
95
-------
Table 9.1 Natural attenuation and mobilization path-
ways for thorium.
Attenuation
Processes
Radioactive
decay for
short-lived
radioisotopes
to n 228T|-i
^y.y . , in,
234Th)
Precipitation
Adsorption
onto aquifer
minerals (clay
minerals, iron/
manganese
oxyhydroxides;
immobile
fni-rric nf
\\J\ 1 1 lo \J\
nofi ira I
I idiu i di
organic
matter)
Mobilization
Processes
Not applicable
Dissolution
due to
changes in
ground-water
chemistry
Desorption
due to
decreasing
pH, increases
in alkalinity or
complexation
by organic
ligands in
ground water;
reductive
dissolution
of iron/
manganese
oxyhydroxides.
Characterization
Approach
Determination of
ground-water velocity
along relevant
transport pathways
and contaminant mass
release rate from source
areas.
Evaluate formation of
hydrous thorium oxide
along relevant transport
pathways based on
calculated chemical
saturation state of ground
water.
Evaluate total adsorption
capacity of aquifer solids
under representative
ground-water chemistry;
chemical extractions to
assess concentrations
of thorium partitioned to
sorbent phases in aquifer
solids along relevant
transport pathways.
Aqueous Measurements
Available methods for the determination of thorium radionu-
clides in ground water are reviewed in USEPA (2006a) and
Hou and Roos (2008). In general, short-lived radioisotopes
(228Th, 234Th) are best determined by radiometric methods,
while long-lived radioisotopes (229Th, 232Th) can also be
measured using mass spectrometric methods. Alpha spec-
trometry is the most common method used for detection of
228Th, 230Th, and 232Th with some form of separation (e.g.,
extraction chromatography) of the thorium radioisotope from
the ground-water matrix conducted prior to analysis (e.g.,
Grate et al., 1999; Tsaia et al., 2008). The use of mass
spectrometry for analysis of longer-lived thorium radioiso-
topes is discussed by Lariviere et al. (2006) and Rozmaric
et al. (2009). As with radiometric techniques, separation
of the thorium radioisotope from the water matrix improves
analysis by concentrating thorium for improved sensitivity
and isolating thorium radioisotopes from potentially interfer-
ing constituents such as major ions and/or other actinides.
Solid Phase Measurements
The types of solid phase measurements that may be
needed to assess thorium attenuation via immobilization
and to differentiate natural versus contamination sources
within the plume include determinations of total thorium
concentrations, activity/concentration of specific thorium
radioisotopes, as well as the partitioning of thorium to
specific solid phase components along relevant transport
pathways. For complete dissolution of the solid matrix, use
of concentrated acid mixtures and/or alkaline fusion to break
down resistant silicate components is typically required (Le
Fevre and Pin, 2002; Selvig et al., 2005; Godoy et al., 2006;
Galindo et al., 2007; Jia et al., 2008). As noted by Blanco
et al. (2005) and Shimada-Fujiwara et al. (2009), use of
hydrofluoric acid may lead to loss of thorium via precipita-
tion of insoluble thorium fluoride during procedures that
employ this strong acid (e.g., SW-846 Method 3052) if steps
are not taken to prevent its formation. As with measure-
ments for ground-water samples, isolation of the dissolved
thorium radioisotopes from the sample matrix is commonly
employed to improve the accuracy of radioisotope deter-
minations. As demonstrated by Ketterer et al. (2000a;
2000b), analysis of the ratio of thorium radioisotopes in solid
samples may assist in differentiating potential sources of
elevated thorium concentrations in aquifer solids. Thorium
contamination derived from the processing of uranium-
bearing materials would likely have distributions of thorium
radioisotopes that differ significantly from that observed
in natural or background settings at a site. For example,
depleted uranium wastes with 225y:23sy ratios much lower
than observed in aquifer solids from background locations
may also have 230Th:232Th ratios higher than observed in
background locations due to 230Th derived from the 238U
decay series. Sims et al. (2008) recently reported on the
results of an inter-laboratory comparison for the determina-
tion of 230Th:232Th ratios in a wide range of solid reference
materials, demonstrating the potential applicability of this
forensic tool for site investigations.
Sequential extraction procedures may be employed to help
identify the reactive solid component(s) controlling thorium
adsorption within the contaminant plume. Determination
of the host mineral phase(es) dissolved for each extrac-
tion step is recommended, along with the use of surrogate
thorium-bearing phases spiked into the sediment to confirm
accuracy of the procedure (e.g., Rudd et al., 1988). The
choice of appropriate thorium surrogate phase(s) would
be governed by site-specific geochemical conditions or
characterization of the mineralogy of the aquifer sediment.
Based on information for likely endpoints for solid-phase
partitioning, thorium adsorbed to existing/newly formed iron/
manganese oxides or clay minerals may provide reason-
able surrogates. Spiking aquifer solids with these surrogate
phases provides the basis for assessing the actual selectiv-
ity of the various extraction solutions used to target specific
solid components. Physical and chemical procedures,
e.g., size fractionation, may also be employed to assist in
identifying specific components within aquifer solids that
dominate thorium solid-phase partitioning. Examples of
96
-------
procedures to identify the type and abundance of specific
minerals along transport pathways are available in the lit-
erature (e.g., Amonette, 2002). Additional information on
analysis approaches and analytical techniques applied to
solid phase characterization is provided in USEPA (2007;
Section 1MB).
Long-term Stability and Capacity
Limitations for the attenuation of thorium via adsorption
onto aquifer solids are not anticipated due to the relative
abundance of reactive surface area for uptake. However,
this behavior is predicated on the absence of dissolved/
mobile constituents that may compete for binding with
thorium. As discussed previously, formation of soluble
complexes with carbonate or dissolved organic carbon
can increase the mobility of thorium. Examples where the
concentrations of these constituents might be anticipated
include 1) contaminant plumes in which recalcitrant and/
or degradable organic contaminants are also present and
2) shallow ground-water systems in which a downward flux
of natural sources of dissolved organic carbon from overly-
ing soils (e.g., humic or fulvic components) is supported
by surface infiltration from surface sources of water and/or
fluctuations of the shallow ground-water table. Recalcitrant
organic contaminants, such as synthetic chelating agents
that may have been used in process solutions, may facili-
tate thorium migration via formation of mobile complexes in
ground water. Degradation of organic contaminants within a
plume will lead to the production of excess alkalinity relative
to the natural ground-water conditions, which could impede
adsorption or precipitation of thorium via complexation
with dissolved carbonate. Also, the spread of reducing
conditions that may accompany down gradient transport
of organic contaminants could lead to reductive-dissolution
of iron/manganese oxyhydroxides and re-mobilization of
adsorbed or coprecipitated thorium.
Due to the tendency for thorium to bind with natural organic
matter, periodic influxes of these components from overlying
soils may lead to expansion of the thorium plume in shallow
ground-water systems (e.g., Marley et al., 1993). However,
results from monitoring of colloid concentrations within
landfill leachate and down gradient monitoring locations of
an impacted aquifer indicate that the distance of colloid-
facilitated transport may be limited (Baumann et al., 2006).
For shallow ground-water plumes, the spatial distribution
of monitoring locations will thus be an important factor in
assessing the impact of natural organic matter migration on
thorium transport. As an example, McCarthy et al. (1998a;
1998b) have described field investigations for a historical
shallow, unlined disposal site for radioactive waste in which
episodic radionuclide leaching from subsurface waste
residuals has occurred due to interactions with surface
water infiltration and/or water table fluctuations that mobilize
natural organic matter from overlying soils. Contaminant
migration occurred along preferential flow paths through a
fractured bedrock system, discharging via down gradient
seeps approximately 80 meters from the shallow disposal
area. For sites with similar disposal scenarios, the physi-
cal characteristics of the disposal system and the shallow
hydrologic system (water flow and chemistry) are critical
factors to evaluate relative to the sustainability of natural
attenuation processes.
Tiered Analysis
Determination of the viability of monitored natural attenu-
ation for remediation of long-lived thorium radioisotopes
in ground water will depend upon proper assessment of
contaminant loading to the aquifer relative to the velocity
of ground-water travel and the prevailing geochemistry
and mineralogy within the contaminant plume and the
down gradient zone prior to the point(s) of compliance. The
goal of site assessment is to demonstrate the process(es)
controlling thorium sequestration onto aquifer solids and
the long-term stability of solid phase thorium as a function
of existing and anticipated ground-water chemistry. The
following tiered analysis structure for site characterization
provides an approach to evaluate candidate sites and define
the potential limitations of MNA as part of a remedy for
ground-water cleanup.
Tier I. Site characterization under Tier I will involve demon-
stration that the ground-water plume is static or shrinking,
has not reached compliance boundaries, and does not
impact existing water supplies. Once this is established
through ground-water characterization, evidence is col-
lected to demonstrate thorium partitioning to aquifer solids
within the plume. If immobilization processes are active
throughout the plume, then there should be an observed
increase in solid phase concentrations within regions of
the plume with higher aqueous concentrations, e.g., near
the source term. Evaluation of the mass/activity of tho-
rium distributed between ground water and aquifer solids
throughout the plume is recommended to account for both
existing and potentially mobile forms of thorium. This field
partitioning data may be supplemented by geochemical
modeling that incorporates measured water chemistry
(e.g., pH and major ion chemistry) throughout the plume to
assess the potential for aqueous thorium species that may
limit precipitation or adsorption of thorium along relevant
transport pathways (e.g., complexation with carbonate or
dissolved organic matter). It is also important at this stage
of the site evaluation to determine source term character-
istics such as the inventory of contaminant mass and the
current and historical rate of release into the subsurface.
Determination of the distribution of thorium radioisotopes
in ground water and aquifer solids both upgradient and
within the plume is recommended to assist in differentiating
between natural and anthropogenic source contributions.
For example, the presence of 229Th and/or 228Th (due to its
short half-life) would indicate a source of thorium other than
natural levels derived from the aquifer. These assessments
could be supplemented by examination of radionuclide
production/process history and projected inventories for
the site. Acquisition of this information in combination with
identification of a stable plume provides justification for
proceeding to Tier II characterization efforts.
Tier II. Under Tier II, the apparent rate and mechanism(s) of
attenuation are determined. Estimates of a site attenuation
97
-------
rate(s) can be assessed via a well transect along the
ground-water flow path. In addition, time-series data may
be collected at one or more monitoring points within the
plume (USEPA, 2007; Section IIIA.5). This information
will allow assessment of the relative timescales for con-
taminant attenuation and fluid transport and determination
of whether remediation objectives can be met within the
required regulatory time frame. If concentrations of short-
lived thorium radioisotopes (e.g., 228Th, 234Th) are elevated
within a plume, it may be important to determine the rela-
tive contributions of radioactive decay and immobilization
to the overall observed attenuation. Determination of the
contribution of radioactive decay will necessitate detailed
analysis of system hydrology relative to flow pathway(s),
flow velocity, and temporal variations in flow velocity and/or
direction within the boundaries of the plume. This informa-
tion, in combination with knowledge of contaminant source
release characteristics, can be employed to develop a
decay-transport model to project the activity/concentration
distribution of thorium radioisotopes throughout the plume in
the absence of adsorption or precipitation processes. For
systems in which immobilization plays a role in observed
attenuation, it will be necessary to identify whether adsorp-
tion onto existing aquifer minerals or newly formed minerals
predominates. Measurements of changes in ground-water
chemistry along a transport pathway, e.g., decreasing fer-
rous iron concentrations coupled to increasing dissolved
oxygen concentrations, may be used to infer formation of
newly formed minerals. Confirmation of thorium adsorp-
tion onto existing or newly formed aquifer minerals will
necessitate determination of the chemical speciation of
solid phase thorium and may be approached according to
the following scheme:
1.
2.
Calculation of saturation state of ground water
relative to precipitation of hydrous thorium oxide
and potential newly formed aquifer minerals (e.g.,
carbonates, Fe/Mn (hydr)oxides) along relevant
thorium transport pathways;
Determination of aquifer mineralogy to determine
the relative abundance of components with
documented capacity for thorium adsorption , with
implementation of steps for aquifer solids collection,
processing and analysis that avoid transformation
of mineral species from reduced zones (e.g.,
oxidation of ferrous carbonate/sulfide to ferric (hydr)
oxide; USEPA, 2006b); and
3.
Determination of thorium-sediment associations
via chemical extractions designed to target specific
components within the aquifer sediment.
This compilation of information will facilitate identification
of the reaction(s) leading to thorium attenuation. The
demonstration of concurrence between conceptual and
mathematical models describing thorium transport will entail
development of site-specific parameterization of the chemi-
cal processes controlling thorium solid phase partitioning.
Tier III. Once the contributions from radioactive decay and
adsorption/precipitation processes have been determined,
the subsequent characterization effort under Tier III will
involve determination of the stability of immobilized tho-
rium and the capacity of the aquifer to sustain continued
uptake. It is recommended that the stability of immobilized
thorium be tested based on the anticipated evolution of
ground-water chemistry concurrent with plume evolution.
For example, changes in ground-water pH and/or alkalin-
ity can exert a significant influence on thorium adsorption.
Therefore, it is recommended that leach tests with aquifer
solids be conducted to characterize the magnitude of tho-
rium re-mobilization as a function of pH for a ground-water
chemistry representative of site conditions. It is recom-
mended that the capacity for thorium uptake onto aquifer
solids be determined relative to the specific mechanism(s)
identified in Tier II. For sites in which a continuing source of
thorium to the saturated aquifer exists, it is recommended
that potential steps to minimize or eliminate this continued
contaminant flux be evaluated and implemented where
feasible. If site-specific tests demonstrate that the stability
and capacity for thorium immobilization, in combination with
continued elimination of short-lived thorium radioisotopes
via radioactive decay, are sufficient to sustain attenuation,
then the site characterization effort can progress to Tier IV.
Tier IV. Finally, under Tier IV a monitoring plan is estab-
lished along with contingency plans in the event of MNA
failure. It is recommended that the monitoring plan be
designed to establish both continued plume stability and to
identify changes in ground-water chemistry that may lead to
re-mobilization of attenuated thorium. The specific chemi-
cal parameters to be monitored will include those identified
under Tier III that may halt thorium partitioning to aquifer
minerals and/or result in solubilization of thorium precipi-
tates. Solution phase parameters that could alter the extent
of thorium adsorption include decreases in pH, increases
in alkalinity, and/or increases in the concentration of DOC
in ground water. Changes in water chemistry may occur
prior to observed changes in dissolved thorium and, thus,
serve as monitoring triggers for potential MNA failure. Sites
at which residual anthropogenic contaminant sources are
left in unsaturated zones should include monitoring points
to assess changes in the release of thorium to the satu-
rated aquifer due to increased surface infiltration or rises
in the ground-water table. Changes in system hydraulics
may serve as monitoring triggers for potential MNA failure.
In this instance, a contingency plan can be implemented
that incorporates engineered strategies to arrest possible
plume expansion beyond compliance boundaries. Possible
strategies to prevent plume expansion include ground-water
extraction with surface treatment, installation of permeable
reactive barriers to enhance uptake capacity perpendicu-
lar to the direction of plume advance, or enhancement of
adsorption capacity within the aquifer through the injection
of soluble reactive components that precipitate within the
formation.
98
-------
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101
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Cesium
Paul M. Bertsch, James E. Amonette, Robert G. Ford
Occurrence and Distribution
As a major fission product, the introduction of radioisotopes
of cesium to the environment is primarily a result of fallout
from nuclear weapons testing (e.g., Robison et al., 2003),
accidental releases from power and nuclear materials pro-
duction reactors (e.g., Korobova et al., 2008), and releases
from radioactive materials processing/storage facilities. To
a much lesser degree, 137Cs can be introduced into the
environment or pose a risk to humans as a result of its use
in a number of industrial devices, such as moisture-density
gauges, widely used in the construction industry, leveling
gauges, used in industries to detect liquid flow in pipes
and tanks, thickness gauges, for measuring thickness of
sheet metal, paper, film and many other products, as well
as well-logging devices used in the drilling industry to help
characterize rock strata. Hospitals and associated research
laboratories are another potential source of 137Cs to the
environment, due to the use of this radioisotope in medical
therapies to treat cancer. Stable cesium (133Cs) is also pres-
ent in the environment at concentrations significantly higher
than typically encountered for its radioactive isotopes (e.g.,
Kreamer et al., 1996; Stetzenbach et al., 1999; Millings et
al., 2003). The largest source of 133Cs is in the form of the
mineral pollucite (CsAISi2O6; Teertstra and Cerny, 1995),
which is recovered during mining for commercial uses of
stable cesium (Butterman et al., 2005).
Geochemistry and Attenuation Processes
Radioactive Decay
Four isotopes of cesium are produced from thermal neutron
fission of 235U, including 133Cs (stable), 134Cs (t1/2 = 2.07
years), 135Cs (t1/2 = 2.3x10s years), and 137Cs (t1/2 = 30.0
years) (Isnard et al., 2009). The stable isotope, 133Cs, pre-
dominates in natural systems that have not been impacted
by releases of fission products generated from controlled/
uncontrolled nuclear reactions and from the production or
processing of nuclear fuels. The radioisotope, 134Cs, is
produced from neutron capture by fission-produced 133Cs,
but only at low yield (Chung et al., 1992). The amount of
135Cs produced during fission is governed by the extent
that its parent, 135Xe, is converted to 136Xe during neutron
irradiation (Hou and Roos, 2008). However, due to its long
half-life, 135Cs may persist at impacted sites even if initially
present at substantially lower levels than 137Cs (Pibida et al.,
2004). The decay of 137Cs to stable 137Ba occurs via both a
direct pathway (5%) and an indirect pathway via metastable
i3/mga (95%). |n general, 137Cs is of greatest concern from
a radiation exposure perspective, since decay of the 137mBa
(t1/2 = 2.552 min) intermediate results in the emission of a
662 keV gamma ray. Some general aspects of the decay
reactions are illustrated in Figure 10.1.
neutron
capture
136Xe
stable
135|
6.61 h
135Xe
9.09 h
135Cs
2.30x1 0"y
135Ba
stable
137Cs
30.0 y
137Ba
stable
137mgg
2.552 m
662keV7
Figure 10.1 Decay reactions involving 13sCs and 137Cs.
Radioactive decay for 137Cs occurs via both a
metastable intermediate (137mBa) and directly
to 137Ba via ejection of a beta particle (£>-). The
decay of 137mBa to 137Ba involves emission of a
662 keV gamma (y) ray. The half-life is shown
directly below the isotope that is subject to
decay; y=years, h=hours, m=minutes, stable
=non-radioactive isotope. Decay half-life data
were obtained using the Win Chain program
that provides electronic access to the ICRP38
Nuclear Decay Data Files (ICRP, 1983;
Eckerman et al., 1994). WinChain is a pub-
lic domain software application available for
download from Oak Ridge National Laboratory
(http://ordose. ornl. gov/down loads, html).
Aqueous Speciation
Cesium is a highly water soluble alkali metal, and, as with
other alkali metals, it exists as a monovalent cation that
forms few precipitation products. Additionally, cesium exists
in the monovalent oxidation state throughout the range of
redox conditions normally found in nature (USEPA, 1999).
Since cesium does not form insoluble precipitates in low-
temperature subsurface systems, adsorption onto aquifer
solids is the dominant process controlling its solid phase
partitioning in ground water.
Adsorption
The primary process for cesium adsorption onto aquifer
solids is via ion exchange onto minerals with layered struc-
tures and high cation exchange capacity, including 2:1 layer-
silicate minerals and layered manganese oxides (USEPA,
1999; Siegel and Bryan, 2003; Lopano et al., 2009). While
solid components such as natural organic matter or variable
103
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charge oxide minerals may possess negatively-charged
surface sites for cation exchange, cesium exchange with
these phases is generally observed to be readily reversible
(Valcke and Cremers, 1994). Cesium is a large cation with
a tendency to easily lose water molecules that coordinate
with this ion in groundwater. When cesium ions migrate into
the interlayer spaces within expandable 2:1 layer-silicates
and lose these waters of hydration, the clay interlayers can
collapse and limit the exchangeability of adsorbed cesium.
Models that successfully predict cesium adsorption data
onto soils typically distribute the total density of adsorption
sites between fractions with low and high selectivity/affinity
for cesium (e.g., Bradbury and Baeyans, 2000; Zachara et
al., 2002). Generally, the portion of the aquifer solids that
show greatest selectivity for cesium adsorption is referred
to as the micaceous fraction.
Specific sorption of cesium to micaceous fractions of
soils is strong with the greatest sorption occurring along
the weathered edges of micas (Tamura, 1963; Sawhney,
1970; Francis and Brinkley, 1976). Micaceous minerals
(biotite, muscovite) are 2:1 layer-silicates containing
permanent negative-charge due to isomorphic substitution
in the crystal structure. The resulting negative charge is
balanced primarily by potassium within the interlayer space.
Vermiculites and smectites are similar in structure, but have
a lower proportion of edge sites versus the true micas
(Sawhney, 1970). Literature on the sorption of cesium in
soils of varying mineralogy is abundant (Coleman et al.,
1963; Tamura, 1963; Sawhney and Frink, 1964; Sawhney,
1965; Sawhney, 1966a; Sawhney, 1966b; Sawhney, 1972).
Micaceous clays can be transformed into illites, vermiculites,
smectites, or kaolinites depending on the amount they have
been physically, chemically, and biologically weathered.
Generally, as the degree or intensity of weathering
increases, mineral abundance changes in the order of:
mica < illite < vermiculite < smectite < kaolinite, indicating
that less weathered soils would generally show greater
capacity for selective adsorption of cesium. Studies of clay
minerals isolated from contaminated soils have confirmed
that micaceous clay minerals selectively concentrate cesium
(e.g., McKinley et al., 2001; Cha et al., 2006).
Site Characterization
Overview
Attenuation of 137Cs might be achieved through radioactive
decay with the influence of adsorption or ion exchange reac-
tions. In general, since cesium adsorption is dominated by
partially reversible ion exchange reactions, it is anticipated
that the primary function of this mechanism would be to
limit the rate of mass transport to allow radioactive decay
to remove sufficient contaminant mass. A list of potential
attenuation processes is provided in Table 10.1. Two factors
that will dictate the adequacy of attenuation via radioac-
tive decay include the rate of water transport and the total
mass and release rate of 137Cs into the subsurface plume.
Table 10.1 Natural attenuation and mobilization path-
ways for cesium.
Attenuation
Processes
Radioactive
decay
Adsorption or
ion exchange
onto aquifer
minerals (clay
minerals,
layered
manganese
oxides)
Mobilization
Processes
Not
applicable
Desorption
decreasing
competition
from major
in ground
reductive
dissolution of
manganese
oxides.
Characterization
Approach
Determination of
ground-water velocity
along relevant
transport pathways
and contaminant
mass release rate
from source areas.
Evaluate total
adsorption capacity
of aquifer solids
under representative
ground-water
chemistry; chemical
extractions to assess
concentrations of
exchangeable 137Cs
in aquifer solids along
relevant transport
pathways.
Aqueous Measurements
Decay of 137Cs results in emission of p-particles with maxi-
mum energies of 514 keV (94.4%) and 1175 keV (5.4%),
and it is accompanied by high-abundance (85.1%) y-ray
emission with energy of 661.7 keV (USEPA, 2006; Hou and
Roos, 2008). Due to the abundance and low self-absorption
of the y-ray emission, gamma spectrometry is the common
radiometric analysis technique applied for quantification of
137Cs in environmental samples. While mass spectrometry
can be used for analysis of 137Cs, the sensitivity of this
approach is inferior to gamma spectrometry. In contrast,
mass spectrometry can be used for quantification of stable
133Cs and long-lived 135Cs. Interference from matrix ions
with similar mass (e.g., stable 135Ba) can be minimized
using off-line ion exchange (Epov et al., 2004) or in-line
chromatographic separation procedures (Evans et al., 2007;
Liezers et al., 2009), although quantification of 135Cs still
may suffer from polyatomic interferences for inductively
coupled plasma-mass spectrometry (e.g., Epov et al., 2003).
Solid Phase Measurements
Solid phase measurements that may provide information
useful to assessing processes controlling 137Cs retardation
within the aquifer and the capacity along relevant transport
pathways include the determination of 137Cs partitioning
to aquifer solids, the cation exchange capacity (CEC) of
aquifer solids, and identification of aquifer solids mineralogy
104
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that may participate in adsorption reactions. Evaluation of
the mass distribution of 137Cs between co-located ground
water and aquifer solids throughout the plume provides an
assessment of the extent that retardation reactions limit
cesium migration. Bulk solid-phase partitioning can be
conducted using total digestion/alkaline decomposition or
acid-extractable techniques (e.g., Taylor et al., 2007; Michel
et al., 2008) with appropriate measures to assess extraction
efficiency (Chang et al., 2004). There are a range of aque-
ous reagent solutions that may be employed to measure
the total CEC of the aquifer solids, as well as the fraction
of exchangeable 137Cs (e.g., Cerling and Spalding, 1982;
Zachara et al., 2002; de Koning et al., 2007). However,
since the measured capacity is dependent on both the
concentration of competing cations and the influence of
ground-water chemistry on the physicochemical charac-
teristics of the exchanging solid phase, a portion of the
tests to determine CEC of aquifer solids should include
water chemistries representative of conditions within the
plume. In addition, tests to assess uptake of 137Cs should
incorporate concentrations of 133Cs that are representative
of site ground water.
An assumption to tests for determination of CEC is that only
exchangeable cations are released from the solid matrix.
It is recommended that the potential for partial dissolution
of the more labile mineral fraction in aquifer solids during
extraction (e.g., Jackson and Inch, 1989) be assessed,
particularly for tests where the extracting solution chemistry
is significantly different from the ground-water chemistry at
locations where aquifer solids are sampled. This may be
achieved through measurement of the major ion chemistry
in the extract solution (e.g., iron, manganese, sulfate, dis-
solved organic carbon), which may serve as markers for the
dissolution/leaching of minerals or organic matter that may
bind 137Cs in a less labile form. Work reported by Bunzl et
al. (1999) indicates that sample drying and storage should
not significantly impact the extractability of cesium in soils.
Physical and chemical procedures, e.g., size fractionation,
may also be employed to assist in identifying specific
components within aquifer solids that dominate cesium
solid-phase partitioning. It should be noted that cesium
partitioning to aquifer solids is not confined primarily to
the clay-sized fraction (<2 |am particle size), so solid-phase
partitioning measurements should include analysis of the
whole aquifer solids sample in addition to any size fractions
that may be isolated from the original sample (Zachara et
al., 2002; Dion et al., 2005; Cha et al., 2006). Examples of
procedures to identify the type and abundance of specific
minerals along transport pathways are available in the
literature (e.g., Cerling and Spalding, 1982; Jackson and
Inch, 1983; McKinley et al., 2001). Additional information
on analysis approaches and analytical techniques applied
to solid phase characterization is provided in USEPA (2007;
Section NIB).
The activity/concentration of 137Cs in shallow aquifers may
be supported, in part, by contributions from historical atmo-
spheric fallout from nuclear weapon testing and infiltration
into the subsurface (e.g., Ciszewski etal., 2008). In general,
these contributions are limited to shallow soils, but the
potential contribution from this source should be recognized
relative to the selection and analysis of aquifer solids for
determination of background concentrations/activities and/
or contaminant distribution relative to source areas. In some
cases, evaluation of the activity ratio of 137Cs and 239+24°Pu
may be used to help differentiate potential sources of 137Cs
in aquifer solids (e.g., Hodge et al., 1996; Kirchner et al.,
2002; Turner et al., 2003). This approach to contaminant
source delineation would require some knowledge of the
ratio of these radioisotopes in potential sources based on
site process history/inventories and/or measurements of
samples collected from source areas.
Long-term Stability and Capacity
The long-term stability of 137Cs attenuated through adsorp-
tion (ion exchange) will depend upon the stability of the
host mineral and the abundance of other ions which might
displace adsorbed cesium. Increases in monovalent cation
levels can be expected to work against 137Cs immobilization
for sites dominated by cation exchange reactions. While
strong binding onto micaceous clay minerals is evident from
laboratory and field studies, there is sufficient evidence
to support that the variability of ground-water chemistry
and aquifer solids properties anticipated at different sites
precludes reliance on immobilization as the attenuation
endpoint (Comans et al., 1989; Smith and Comans, 1996; de
Koning and Comans, 2004; Pinder III et al., 2005). Review
of the extent of plume development for a number of sites
indicates that cation exchange within the saturated aquifer
may have insufficient stability to prevent plume expansion
(Brady et al., 2002). Thus, a critical factor for assessing
the overall capacity of the aquifer for attenuation will be
evaluation of the mass flux of 137Cs moving through the
plume relative to the rate of water movement through the
aquifer. Field studies have indicated that 137Cs in contami-
nated soils can migrate down into shallow ground water as
a result of infiltration of surface water sources (Pinder III
et al., 2005; Paller et al., 2008). Thus, in order to make a
reliable assessment of the mass/activity flux of 137Cs into the
plume, it will be important to understand the characteristics
of the hydrogeologic system and the dynamics of water and
contaminant transfer from contaminant source areas into
the plume. Radioactive decay may be sufficient to prevent
plume expansion, but this is not likely for sites with an
uncontrolled source of 137Cs entering the subsurface and/
or characteristic times for ground-water transport that are
significantly shorter than the half-life of radioactive decay.
Tiered Analysis
Determination of the viability of 137Cs remediation in ground
water via monitored natural attenuation will depend upon
proper assessment of contaminant loading to the aquifer
relative to the velocity of ground-water travel and the
prevailing geochemistry and mineralogy within the con-
taminant plume and the down gradient zone prior to the
point(s) of compliance. The goal of site assessment is to
demonstrate that retardation of 137Cs migration within the
plume is adequate to allow radioactive decay to reduce
105
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137Cs mass/activity to below required cleanup levels. The
following tiered analysis structure for site characterization
provides an approach to evaluate candidate sites and define
the potential limitations of MNA as part of a remedy for
ground-water cleanup.
Tier I. Site characterization under Tier I will involve demon-
stration that the ground-water plume is static or shrinking,
has not reached compliance boundaries, and does not
impact existing water supplies. Once this is established
through ground-water characterization, evidence is col-
lected to demonstrate 137Cs partitioning to aquifer solids
within the plume. If ion exchange processes are active
throughout the plume, then there should be an observed
increase in solid phase concentrations within regions of the
plume with higher aqueous concentrations, e.g., near the
source term. Evaluation of the mass/activity of 137Cs distrib-
uted between ground water and aquifer solids throughout
the plume is recommended to account for both existing
and potentially mobile forms of 137Cs. It is also important
at this stage of the site evaluation to determine source
term characteristics such as the inventory of contaminant
mass and the current and historical rate of release into the
subsurface. Acquisition of this information in combination
with identification of a stable plume provides justification
for proceeding to Tier II characterization efforts.
Tier II. Under Tier II, the apparent rate and mechanism(s)
of attenuation are determined. Estimates of a site attenu-
ation rate(s) can be assessed via a well transect along the
ground-water flow path. In addition, time-series data may
be collected at one or more monitoring points within the
plume (USEPA, 2007; Section IIIA.5). This information will
allow assessment of the relative timescales for contaminant
attenuation and fluid transport and determination of whether
remediation objectives can be met within the required
regulatory time frame. As part of this analysis, it is critical
to determine the relative contributions of radioactive decay
and immobilization to the overall observed attenuation.
Determination of the contribution of radioactive decay will
necessitate detailed analysis of system hydrology relative
to flow pathway(s), flow velocity, and temporal variations in
flow velocity and/or direction within the boundaries of the
plume. This information, in combination with knowledge
of contaminant source release characteristics, can be
employed to develop a decay-transport model to project
137Cs activity/concentration distribution throughout the
plume in the absence of adsorption. Assessment of the
impact of adsorption on the rate of 137Cs migration will
necessitate determination of the extent of solid phase
partitioning along transport pathways within the plume
according to the following scheme:
1. Determination of aquifer mineralogy to determine
the relative abundance of components with
documented capacity for cesium adsorption (e.g.,
Jackson and Inch, 1983; Jackson and Inch, 1989;
McKinley et al., 2001; Amonette, 2002), and
2. Determination of 137Cs -sediment associations via
chemical extractions designed to target specific
components within the aquifer sediment (e.g.,
Cerling and Spalding, 1982; McKinley et al., 2001;
Zachara et al., 2002).
This compilation of information will facilitate identification of
the reaction(s) leading to 137Cs attenuation. The demonstra-
tion of concurrence between conceptual and mathematical
models describing cesium transport (including 133Cs) will
entail development of site-specific parameterization of
the chemical processes controlling cesium solid-phase
partitioning.
Tier III. Once the contributions from radioactive decay
and adsorption (ion exchange) processes has been deter-
mined, the subsequent characterization effort under Tier
III will involve determination of the stability of immobilized
137Cs and the capacity of the aquifer to sustain retardation
of 137Cs transport. It is recommended that the stability
of immobilized 137Cs be tested based on the anticipated
evolution of ground-water chemistry concurrent with plume
evolution. For example, changes in ground-water pH and/or
cation composition can exert a significant influence on 137Cs
adsorption. Therefore, it is recommended that sediment
leach tests be conducted to characterize the magnitude of
137Cs re-mobilization as a function of pH for a ground-water
chemistry representative of site conditions (including 133Cs
in site ground water). It is recommended that the capacity
for 137Cs uptake onto aquifer solids be determined rela-
tive to the specific mechanism(s) identified in Tier II. For
sites in which a continuing source of 137Cs to the saturated
aquifer exists, it is recommended that potential steps to
minimize or eliminate this continued contaminant flux be
evaluated and implemented where feasible. If site-specific
tests demonstrate that the stability and capacity for 137Cs
adsorption, in combination with continued elimination of
137Cs via radioactive decay, are sufficient to sustain plume
attenuation, then the site characterization effort can prog-
ress to Tier IV.
Tier IV. Finally, under Tier IV a monitoring plan is estab-
lished along with contingency plans in the event of MNA
failure. It is recommended that the monitoring plan be
designed to establish both continued plume stability and
to identify changes in ground-water chemistry that may
lead to re-mobilization of attenuated cesium (both 137Cs
and stable 133Cs). The specific chemical parameters to be
monitored will include those identified under Tier III that may
halt cesium partitioning to aquifer minerals. Solution phase
parameters that could alter cesium adsorption include
changes in pH or increases in the concentrations of com-
peting cations in ground water. As an example, increases
in potassium (K+) or ammonium (NH4+) concentrations in
ground water could signal the potential for displacement of
137Cs from cation exchange sites. Changes in water chem-
istry may occur prior to observed changes in solution 137Cs
and, thus, serve as monitoring triggers for potential MNA
failure. In this instance, a contingency plan can be imple-
mented that incorporates engineered strategies to arrest
possible plume expansion beyond compliance boundaries.
Possible strategies to prevent plume expansion include
ground-water extraction with surface treatment or instal-
lation of permeable reactive barriers to enhance uptake
106
-------
capacity perpendicular to the direction of plume advance.
For sites in which residual subsurface contamination is left
in place, it is also recommended that the monitoring plan
be designed to identify changes in ground-water levels, flow
velocity, or direction that might influence the efficiency of
137Cs removal via radioactive decay. In particular, sites at
which residual 137Cs sources are left in unsaturated zones
should include monitoring points to assess changes in the
release of 137Cs to the saturated aquifer due to increased
surface infiltration or rises in the ground-water table.
Changes in system hydraulics may serve as monitoring
triggers for potential MNA failure.
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109
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110
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Plutonium and Americium
Patrick V. Brady, Robert G. Ford, Richard T. Wilkin
Occurrence and Distribution
Elevated concentrations of the radioisotopes of plutonium
and americium in ground water are generally attributed to
man-made sources such as the production and process-
ing of nuclear materials for power generation or weapons
production and testing. Trace amounts of 239Pu exist in
nature, primarily in uranium ore deposits, as a result of
neutron capture by 238U (Curtis et al., 1999; Wilcken et al.,
2008). Low levels of plutonium radioisotopes are distrib-
uted throughout surface soils due to historical atmospheric
testing of nuclear weapons, and the radioisotopes 240Pu,
241 Pu, and 242Pu are considered to be derived solely from
anthropogenic sources (Ketterer and Szechenyi, 2008).
Releases of plutonium and americium radioisotopes from
shallow, subsurface disposal systems (Carlton, 1997;
Groenewold et al., 2005; Cantrell, 2009) and/or accidental
releases from processing or storage facilities for nuclear
materials (e.g., Lee and Clark, 2005) account for the major-
ity of elevated concentrations observed in the subsurface
at federal or commercial sites across the United States (Hu
et al., 2010). The primary source of 241Am in the environ-
ment is derived from the production and disposal/release
of plutonium. As discussed below, these radionuclides
are typically encountered together due to parent-daughter
relationships in relevant decay chains.
Geochemistry and Attenuation Processes
Radioactive Decay
There are fifteen radioisotopes of plutonium; the three most
common ones are 239Pu (-94%), 240Pu (-6%), 241Pu (0.4%).
The decay half-life of 239Pu is 24,100 years, and it decays
to 235U with the emission of an alpha particle (Figure 11.1).
The decay half-life of 240Pu is 6540 years, producing 236U
via alpha decay (Figure 11.2). The 241Pu radioisotope
(t1/2= 14.4 years) decays to241 Am (t1/2 = 432 years) via emis-
sion of a beta particle (Figure 11.3). The 241Am radioisotope
subsequently decays to 237Np (t1/2 = 2.14 million years) with
the emission of an alpha particle. Possible progenitors
that could introduce uranium isotopes via decay in-growth
are shown in Table 11.1. Transuranic radioisotopes such
as americium, curium, and neptunium decay to produce
plutonium and/or americium radioisotopes of importance
(e.g., Carlton, 1997; Dai et al., 2002). The decay half-
lives for these progenitors range from minutes to many
thousands of years. Knowledge of the presence of these
radionuclides within the plume and sampled ground water
may be important for proper identification of the source of
plutonium/americium radioisotopes as well as in-growth
corrections that may be required to properly account for
the mass/activity of these radioisotopes at the time of
sample collection.
Table 11.1 Illustration of potential decay paths from
different progenitor sources leading to
production of americium and plutonium
radioisotopes. Determination of possible
decay paths to the target radionuclide
was based on examination of the Chart
of Nuclides (http://www.nndc.bnl.gov/
chart/) maintained by the Brookhaven
National Laboratory, National Nuclear Data
Center (NNDC) relative to possible decay
paths based on decay modes identified
in the Appendix (EC = electron capture,
p- = electron emission, a = alpha decay).
Decay half-life data were obtained using the
WinChain program that provides electronic
access to the ICRP38 Nuclear Decay
Data Files (ICRP, 1983; Eckerman et al.,
1994; m = minutes, h = hours, d = days,
y = years). WinChain is a public domain
software application available for download
from Oak Ridge National Laboratory
(http://ordose. ornl. gov/dow nloads.html).
Contaminant
Radionuclide
241 Am
239Pu
240pu
241 Pu
Decay
Progenitor
241 Cm
241 Pu
239Np
243Cm
243Am
(via 239Np)
240Am
240Np
244Cm
245Cm
Decay
Mode
EC
P-
P-
a
a(p-)
EC
P-
a
a
Progenitor
Decay
Half-life
32.8 d
14.4 y
2.355 d
28.5 y
7380 y
(2.355 d)
50.8 h
65 m
18.11 y
8500 y
111
-------
239pu
Decay
Series
6.54x103y
a
240pu
Decay
Series
236(J
2.34x107y
a
1.40x1010y
a
228Ra
5.75 y
228Ac
6.13h
a
224Ra
3.66d
a
220Rn
55.6s
a
a
212Bj
60.5m
a
208T|
3.07m
(0.3594)
o.3ns
(0.6406)
a
7Pb
Figure 11.1 Decay series for 239Pu. The half-life is
shown directly below the isotope that is
subject to decay; y = years, d = days, h =
hours, m = minutes, s = seconds, stable =
non-radioactive isotope. Decay half-life data
were obtained using the WinChain program
that provides electronic access to the ICRP38
Nuclear Decay Data Files (ICRP, 1983;
Eckerman et al., 1994). WinChain is a public
domain software application available for
download from Oak Ridge National Laboratory
(http://ordose. ornl. gov/downloads. html).
Figure 11.2 Decay series for 240Pu. The half-life is
shown directly below the isotope that is sub-
ject to decay; y = years, h = hours, d = days,
m = minutes, s = seconds, |as = microseconds,
stable = non-radioactive isotope. Decay half-
life data were obtained using the WinChain
program that provides electronic access to
the ICRP38 Nuclear Decay Data Files (ICRP,
1983; Eckerman et al., 1994). WinChain is a
public domain software application available for
download from Oak Ridge National Laboratory
(http://ordose. ornl. gov/downloads.html).
112
-------
241 pu
14.4 y
241 Am
432 y
a
241 pu
Decay
Series
237Np
2.14x106y
a
233pg
27. Od
233|J
1.58x105y
a
229Th
7.34x103y
a
225Ra
14.80d
225Ac
10.0d
a
221 p,-
4.80m
a
217At
32.3ms
a
213B|
45.59m
4.2ns
a
209pb
3.25 h
209B|
stable
Figure 11.3 Decay series for241Pu, including daughter
241 Am. The half-life is shown directly below
the isotope that is subject to decay; y = years,
h = hours, m = minutes, ms = milliseconds,
(is = microseconds, stable = non-radioactive
isotope. Decay half-life data were obtained
using the WinChain program that provides
electronic access to the ICRP38 Nuclear
Decay Data Files (ICRP, 1983; Eckerman
et al., 1994). WinChain is a public domain
software application available for down-
load from Oak Ridge National Laboratory
(http://ordose. ornl. gov/downloads. html).
Aqueous Speciation
Plutonium has four possible oxidation states (i.e., +3, +4,
+5, and +6) under typical subsurface conditions and can
exist in any of these in aqueous solutions. In general, Pu(IV)
species are considered to be most common, but all oxida-
tions may play a role in controlling plutonium speciation
and migration (e.g., Clark et al., 1995). Comprehensive
reviews on the chemical thermodynamics of plutonium
include those by Lemire et al. (2001) and Guillamont et al.
(2003), which address the various aqueous species likely
to be encountered under typical ground-water conditions.
Examples of plutonium species for these four oxidation
states are illustrated in Figure 11.4, including Pu(lll)3+,
Pu(IV)O2, Pu(V)O2+, and Pu(VI)O2(CO3)22-. The distribution
of plutonium species between reduced [Pu(lll), Pu(IV)] and
oxidized [Pu(V), Pu(VI)] forms is considered to exert the
greatest influence on the mobility of this radionuclide in
ground water (e.g., Dai et al., 2002; Kaplan et al., 2004).
Plutonium may form soluble complexes with a range of
anions common to ground water, and these reactions can
play a role in the predominant plutonium oxidation state
(Choppin et al., 1997). Tetravalent plutonium [Pu(IV)] forms
hydrolysis and/or polymeric species in water to a greater
extent than the other oxidation states (Baes and Mesmer,
1976), although these reactions may contribute to Pu(VI)
speciation at elevated concentrations encountered near
source zones (e.g., Reilly and Neu, 2006). Plutonium in its
Pu(VI), Pu(V), and Pu(IV) oxidation states may also form
dissolved complexes with carbonate (Clark et al., 1995;
Clark etal., 1998; Topinetal., 2009). Formation of solution
complexes with synthetic organic chelation agents present
in co-contaminants may also occur, which has motivated
the review of available thermochemical data for the purpose
of calculating potential impacts on plutonium chemical
speciation (Hummel et al., 2005). While the formation
solution complexes with synthetic ligands (e.g., EDTA) is
supported within the technical literature (Meyer et al., 2007;
Hummel et al., 2007), the relative impact of these species
on plutonium transport will be limited by competition reac-
tions with dissolved and solid constituents along relevant
flow paths within the aquifer. As demonstrated by Rai et
al. (2008), the chemical speciation of strong complexants
such as EDTA may be dominated by reactions with aquifer
constituents that are present at much higher concentra-
tions than plutonium. For shallow ground-water systems
with elevated concentrations of natural dissolved organic
carbon, plutonium may form soluble complexes with dis-
solved humic/fulvic compounds (e.g., Marquardt et al., 2004;
Reiller, 2005; Dardenne et al., 2009; Szabo et al., 2010).
The chemistry of americium is less complex than that of
plutonium, since americium exists primarily in the trivalent
state under natural conditions (Clark et al., 1995). The
aqueous chemistry of Am(lll) has been reviewed in Silva et
al. (1995) and updated in Guillamont et al. (2003). As with
plutonium, formation of complexes with dissolved carbonate
can be important in ground water with elevated alkalinity
(e.g., Clark etal., 1995; Vercouter etal.,2005). As shown in
Figure 11.5, carbonate complexes can dominate americium
speciation in solution. Americium may also form complexes
113
-------
with humic/fulvic compounds (e.g., Kim etal., 1993). It has
been demonstrated that formation of soluble complexes with
humic/fulvic compounds can potentially facilitate transport
of Am(lll) in shallow ground water where surface recharge
events can result in increased organic carbon concentra-
tions and aquifer flow velocities (e.g., Marley et al., 1993;
Artinger et al., 1998; McCarthy et al., 1998).
Solubility
Both Am(lll) and Pu(IV) form hydrous oxide phases with low
solubility (Silva et al., 1995; Guillamont et al., 2003; Hala
and Miyamoto, 2007), and these precipitates might control
aqueous concentration of Am(lll) or Pu(IV) near source
zones where elevated concentrations of these radionuclides
might be encountered. As shown in Figure 11.4, precipi-
tation of Pu(IV) phases is projected to dominate over the
other plutonium oxidation states under most ground-water
conditions. However, this also points to the importance
of ground-water redox on the precipitation and stability of
Pu(IV) hydrous oxides. Reduction of Pu(IV) in the presence
of reducing agents such as ferrous iron could potentially
increase overall plutonium solubility (e.g., Rai et al., 2002).
Likewise, oxidation of Pu(IV) could also increase plutonium
solubility. Laboratory solubility studies using systems where
redox and oxygen concentrations have been controlled indi-
cate that hydrous Pu(IV)O2 can be partially oxidized, with
soluble plutonium being controlled by Pu(V) species (Neck
et al., 2007). There is recent evidence that, in the presence
of phosphate, Pu(VI) may behave analogous to U(VI) and
form Pu(VI)-phosphates (Rai et al., 2005). However, there
is currently insufficient knowledge on the likely formation of
this type of precipitate to assess the impact of phosphate
in ground water on aqueous plutonium concentrations in
oxidizing systems. As shown in Figure 11.5, formation of
americium hydroxycarbonate may dominate over hydrous
oxides in ground water with near-neutral pH and elevated
alkalinity. This is supported by laboratory thermochemi-
cal studies that predict hydroxycarbonates as the more
stable phase for actinium(lll) and lanthanum(lll) metals
(Merli et al., 1997). In addition, it has been shown that
Am(lll) readily substitutes for Ca(ll) in the calcite structure
(Curti, 1999; Stumpf et al., 2006). Thus, precipitation of
carbonates may also exert control on americium solubility
in ground water with elevated alkalinity. In general, forma-
tion of solution complexes with synthetic or natural organic
ligands is anticipated to interfere with precipitation of the
various precipitate phases discussed above for plutonium
and americium.
Adsorption
The adsorption of plutonium and americium is anticipated to
be the predominant mechanism controlling solid partitioning
in ground-water systems. A general overview of adsorp-
tion processes for plutonium and americium onto aquifer
solids is provided in USEPA (1999) and USEPA (2004),
respectively. Bradbury and Baeyans (2009) and Degueldre
and Bolek (2009) have demonstrated that americium and
plutonium adsorption onto common aquifer minerals can be
modeled using surface complexation reactions that account
for the influence of pH and competing aqueous speciation
reactions. For americium, formation of soluble complexes
is the predominant factor limiting adsorption to aquifer sol-
ids. However, this influence may be transient depending
on aquifer conditions. For example, while complexation
with soluble humic/fulvic compounds have been observed
to suppress americium adsorption in model systems (e.g.,
Pathak and Choppin, 2007), long-term column studies
have shown that this effect may be limited by competitive
adsorption reactions along the transport pathway (Artinger
et al., 2002). Formation of soluble complexes may also
limit plutonium adsorption. For example, Sanchez et al.
(1985) demonstrated that formation of soluble carbonate
complexes can suppress adsorption of Pu(IV) and Pu(V)
hydrous iron oxides. However, as shown via detailed field
characterization, it appears that controls on the plutonium
oxidation state will exert the greatest influence on plutonium
migration (Dai et al., 2002; Buesseler et al., 2009). For
systems in which more oxidized (and mobile) plutonium spe-
cies have been introduced into the subsurface, interaction
with natural reducing compounds may increase plutonium
adsorption along the flow path through reduction to Pu(IV)
(e.g., Powell et al., 2005; Buerger et al., 2007; Roberts et
al., 2008).
Site Characterization
Overview
Plutonium and americium mobility in ground water is gov-
erned by their total dissolved concentration, the distribution
of plutonium/americium species in water, and the nature
of plutonium/americium partitioning to aquifer solids. The
development of a conceptual site model for predicting the
long-term fate of plutonium/americium at a contaminated
site will require information on the concentration and
chemical speciation of plutonium/americium in the aqueous
phase and the solid phase. Table 11.2 illustrates possible
attenuation and mobilization pathways for plutonium and
americium in ground water. Details of the types of ana-
lytical measurements that may be conducted on sampled
ground water and aquifer sediments to assist in identifying
the attenuation mechanism(s) are discussed in the follow-
ing paragraphs.
Aqueous Measurements
Overviews of radiometric techniques for determining the
activity of plutonium and americium radioisotopes is pro-
vided in USEPA (2006). More recent reviews of the use
of radiometric and mass spectrometric methods for the
quantification of americium and plutonium radioisotopes
is provided in Hou and Roos (2008), Qiao et al. (2009),
and Vajda and Kim (201 Oa; 201 Ob). Of the radioisotopes
of plutonium and americium discussed in this chapter, all
are alpha-emitters except for 241Pu (beta-emitter). Liquid
scintillation counting can be used to measure 241Pu directly
or indirect detection of its daughter, 241Am, can be employed
using alpha spectrometry. The energies of alpha particles
released from 239Pu and 240Pu are too similar to be resolved
using alpha spectrometry (5.147 MeV and 5.170 MeV,
respectively), so the activity of these isotopes is usually
114
-------
5.2
m o
-.2
-.4
-.6
25°C
PuOhT
.2
LU 0
-.2
-.4
-6
25°C
2345678
PH
10 11 12
3 4 5 6 7 8 9 10 11 12
pH
Figure 11.4 Eh-pH stability diagram for Pu at 25 °C and PCO2 = 1Q-25 atm. Diagrams are for solubility control by
PuO2 and Pu(OH)4 (see text). Examples of the four oxidation states of plutonium include Pu(lll)3+, Pu(IV)
02, Pu(V)0+ and Pu(VI)02(C03)/-.
_3
-4
-5
-6
A
I-
OJ
O) -8
_0
-9
-10
-11
-12
Am3
AmOHCO3(c)
O
O
o
o
¥
Am(C03)33-
25°C
10 11 12
PH
Figure 11.5 Solubility diagram forAm(lll) at PCO2 = ia2s atm and PO2 = 0.2 atm.
115
-------
Table 11.2 Natural attenuation and mobilization pathways for plutonium and americium.
Attenuation Processes
Mobilizaton Processes
Characterization Approach
Plutonium
Radioactive decay of
241
Not applicable
Determination of ground-water velocity along relevant
transport pathways and contaminant mass release rate
from source areas.
Reduction of Pu(VI) or
Pu(V) and precipitation
of Pu(IV) hydrous oxide
minerals
Dissolution due to decreased
pH, increased alkalinity, or
oxidation of Pu(IV) to more
mobile species of Pu(VI)/
Pu(V) or further reduction of
Pu(IV) to Pu(lll).
Evaluation of Pu concentration and oxidation state in
ground water and in solid matrix. Evaluation of Pu
solid-phase partitioning using extraction methodologies
coupled to methods to determine Pu oxidation state.
Characterization of aqueous redox and chemical
conditions in ground water with speciation model
evaluation of potential Pu(IV) stability.
Adsorption or
coprecipitation of Pu
with iron/manganese
oxyhydroxides, iron
sulfides, and carbonates
or adsorption onto clay
mineral surfaces
Desorption due to decreased
pH, increased alkalinity, or
high DOC concentrations.
Reductive dissolution of iron/
manganese oxyhydroxides
or oxidative dissolution of
iron sulfides. Oxidation of
adsorbed Pu(IV) to Pu(V)/
Pu(VI).
Evaluation of Pu concentration and oxidation state
in ground water and aquifer solids. Evaluation of Pu
solid-phase partitioning using extraction methodologies;
examine correlation to extractable Fe/Mn, Ca, Mg and
S. Batch and column testing to determine Pu uptake
behavior and capacity of site-specific aquifer materials
under variable geochemical conditions.
Americium
Adsorption or
coprecipitation of Am(lll)
with iron/manganese
oxyhydroxides, iron
sulfides, and carbonates
or adsorption onto clay
mineral surfaces
Desorption due to
decreased pH or high
DOC concentrations (e.g.,
humic/fulvic compounds).
Reductive dissolution of iron/
manganese oxyhydroxides or
oxidative dissolution of iron
sulfides.
Evaluation of Am concentration in ground water and in
solid matrix. Evaluation of Am solid-phase partitioning
using extraction methodologies; examine correlation
to extractable Fe/Mn, Ca, Mg and S. Batch and
column testing to determine Am uptake behavior and
capacity of site-specific aquifer materials under variable
geochemical conditions.
reported as the sum of their combined contribution to
the spectral peak, i.e., 239+24°Pu. The sensitivity of these
methods is generally good for these radioisotopes, but typi-
cally isolation of the analyte from the sample matrix and
interfering radionuclides is required prior to analysis. As
noted by Singhal et al. (2008), high concentrations of natural
organic matter may significantly decrease the efficiency of
target radionuclide isolation. These authors observed that
oxidative destruction of dissolved organic carbon prior to
the separation procedure improved recovery. The yield of
isolation procedures is typically assessed by introducing a
different isotope of the radionuclide that is not anticipated
to be present or detected in the original sample (e.g., Pike
etal.,2009).
Determination of the ratios of plutonium isotopes may be
useful in determining potential contaminant sources (e.g.,
atmosphere fall-out versus site-specific wastes). While
radiometric methods can be employed for this purpose, the
inability to separately determine the activities of 239Pu and
240Pu limits their applicability. Use of mass spectrometric
techniques, including inductively coupled plasma mass
spectrometry (ICP-MS), provides the means to directly
quantify the individual plutonium radioisotopes that can
be used as a signature to differentiate contributions from
different contaminant sources. Reviews of the applicability
and limitations of ICP-MS for quantification of plutonium
radioisotopes can be found in Zoriy et al. (2005), Lariviere
et al. (2006), Hou and Roos (2008), and Ketterer and
116
-------
Szechenyi (2008). Examples where plutonium radioiso-
tope ratios were employed to differentiate contaminant
sources at contaminated sites include Marty et al. (1997;
Los Alamos), Dai et al. (2002; Savannah River Site), and
Dai et al. (2005; Hanford Site). These comparisons are
based on historical or site-specific records of waste char-
acteristics from the production and processing of nuclear
materials, where radioisotope ratios such as 240Pu/239Pu can
be attributed to distinct sources (e.g., 240Pu/239Pu = 0.18
for atmospheric fallout, 0.01-0.07 for weapon production,
or 0.40 for Chernobyl fallout; Warneke et al., 2003). Most
of these studies are based on field sampling campaigns
in which mass spectrometric methods were incorporated
up-front in the field sampling and laboratory analysis plans.
Michel et al. (2007) demonstrated that one can acid-extract
sample planchets prepared for alpha-spectrometry for the
purpose of determining individual 239Pu and 240Pu activities
by ICP-MS. Thus, archived planchet deposits from radio-
metric analyses for a given site may be used to supplement
knowledge of plume characteristics and contaminant source
apportionment.
Determination of the oxidation state(s) of plutonium in
sampled ground water can provide useful information on
the mechanisms controlling transport (e.g., Dai et al., 2002;
Buesseler et al., 2009). Radiochemical methods have
been developed for separating plutonium into its more
reduced [Pu(lll) and Pu(IV)] and more oxidized [Pu(V) and
Pu(VI)] states (Lovett and Nelson, 1981). The accuracy of
these methods is assessed by spiking samples with tracer
radioisotopes with known oxidation state (e.g., 244Pu(lll/IV)
in NIST SRM-996 Spike Assay and Isotopic Standard and
242Pu(V/VI) in SRM-4334C). Description and application of
methods to measure the distribution of plutonium oxidation
states in aqueous samples is provided in several studies in
which species detection is achieved either off-line or in-line
with the detection system (e.g., Grate and Egorov, 1998;
Kuczewski et al., 2003; Powell et al., 2005; Fajardo et al.,
2008). The separation of plutonium species with different
oxidation states is possible due to their inherent differ-
ences in chemical reactivity. In principal, detection of the
separated species can be achieved with either radiometric
or mass spectrometric methods. One advantage of using
mass spectrometry is the ability of directly determining the
distribution of plutonium radioisotopes across the range of
oxidation-state species within the sample. This approach
has proven useful in differentiating the sources of plutonium
contamination sources in ground water at large and/or
complex waste sites (e.g., Buesseler et al., 2009).
Solid Phase Measurements
Analysis of plutonium association with aquifer solids may
be necessary in order to confirm active attenuation of con-
taminants from the ground-water plume and to identify the
reaction process(es) controlling plutonium immobilization.
Determination of the form in which plutonium and americium
is being immobilized underpins projections of the capacity
for continued attenuation of contaminant mass within the
plume and the long-term stability of immobilized radionu-
clides with long decay half-lives. Methods for the extraction
of plutonium/americium from solids are well-documented
(e.g., Croudace et al., 1998; Epov et al., 2007; Varga et al.,
2007; Michel et al., 2008; Payne et al., 2008; Choiniere et
al., 2009; Eikenberg et al., 2009). Since these methods
also extract other actinides or fission products, as well as
matrix elements, isolation of the plutonium/americium radio-
isotopes is a critical step to insure measurement accuracy.
For determination of total solid-phase plutonium, it may also
be necessary to assess the distribution of radioisotopes to
confirm that the extracted radioisotopes are representative
of those being attenuated from the plume.
Sequential extractions using solutions designed to attack
specific solid phase components have been used in an
attempt to better define the chemical speciation of pluto-
nium and americium in solid samples (e.g., Schultz et al.,
1998). This work has demonstrated the difficulty of uniquely
identifying solid phase speciation for plutonium, due to its
tendency to re-adsorb to un-extracted solid components
prior to solid-liquid separation. Schultz et al. (2000) and
Lucey et al. (2007) discuss results on studies designed
to minimize plutonium re-adsorption through addition of a
complexing ligand in the extractant solutions. This work
demonstrates significant improvement in eliminating this
analytical artifact, but these results also show that some
selectivity of an extractant to target a specific solid-phase
association may be lost. In addition, Lucey et al. (2007)
demonstrate the need to take precautions to preserve redox
characteristics of solid samples retrieved from reduced
systems, which dictates application of extractions in an
oxygen-free analytical system (e.g., anaerobic glove box
or closed fluid handling systems). Evaluations have not
been conducted to determine whether these extraction
procedures may influence the in-situ oxidation state of
solid-phase plutonium prior to extraction, so reliance on
spectroscopic methods to identify in-situ plutonium oxidation
state and/or chemical associations is currently warranted
(e.g., Duff et al., 2001; Kaplan et al., 2007).
Long-term Stability and Capacity
The long-term stability of immobilized plutonium will depend
on the maintenance of either 1) sufficiently low reduction
potentials to prevent oxidation and consequent solubiliza-
tion of Pu(IV) solids or 2) stability of the sorbent mineral
and sufficiently low concentrations of competing ions that
could displace adsorbed plutonium. Once plutonium has
been precipitated or adsorbed, the sustainability of the
geochemical driving force (e.g., redox, pH, and/or avail-
able surface sites) is critical to whether natural attenuation
will be a viable cleanup option. The long-term stability of
immobilized americium will depend on the maintenance of
either 1) near-neutral pH and/or sufficiently high alkalinity or
2) stability of the sorbent mineral and sufficiently low con-
centrations of competing ions that could displace adsorbed
americium. Thus, it is recommended that post-attenuation
changes in water chemistry be carefully considered to
ensure that re-mobilization of attenuated plutonium/ameri-
cium does not occur. For plutonium, a particular concern
are situations in which plutonium is attenuated under
reducing conditions that are induced by characteristics of
117
-------
the contaminant plume, specifically if the natural conditions
within the aquifer are more oxidizing. As shown by Dai et
al. (2002) and Buesseler et al. (2009), the more oxidized
forms of plutonium tend to be the most mobile.
Determination of capacity for attenuation will depend on
knowledge of the specific mechanisms leading to plu-
tonium/americium partitioning to aquifer solids and the
flux of these contaminants being transmitted through the
aquifer. Site conditions under which attenuation capacity
may be limited include those in which periodic surface
recharge events flush these contaminants from un-lined,
shallow disposal units where leachate collection systems
do not exist. Factors that tend to interfere with contaminant
immobilization under these settings include ground-water
flow rates that limit the time of contact with aquifer solids
and the introduction of soil-derived organic carbon that
complexes plutonium/americium. The affective capacity for
attenuation within the aquifer will also depend strongly on
the characteristics and variability of ground-water chemistry
and aquifer solids properties along transport pathways, as
well as the impact of hydrologic dynamics on subsurface
chemistry as a function of space and time.
Tiered Analysis
Determination of the viability of plutonium and americium
remediation in ground water via monitored natural attenu-
ation will depend upon proper assessment of contaminant
loading to the aquifer and prevailing geochemistry and
mineralogy within the contaminant plume and the down
gradient zone prior to the point(s) of compliance. While
radioactive decay will contribute to attenuation of 241Pu and
241 Am, decay will not provide a viable mechanism for plume
attenuation for the longer-lived 239Pu and 240Pu. Therefore,
the goal of site assessment will be to demonstrate the
process(es) controlling contaminant sequestration onto
aquifer solids and the long-term stability of solid phase
plutonium and/or americium as a function of existing and
anticipated ground-water chemistry. The following tiered
analysis structure for site characterization provides a techni-
cally defensible approach to evaluate candidate sites and
define the potential limitations of MNA as part of a remedy
for ground-water cleanup.
Tier I. Site characterization under Tier I will involve demon-
stration that the ground-water plume is static or shrinking,
has not reached compliance boundaries, and does not
impact existing water supplies. Once this is established
through ground-water characterization, evidence is col-
lected to demonstrate plutonium/americium partitioning to
aquifer solids within the plume. If natural attenuation pro-
cesses are active throughout the plume, then there should
be an observed increase in solid phase concentrations
within regions of the plume with higher aqueous concen-
trations, e.g., near the source term. This field partitioning
data may be supplemented by geochemical modeling that
incorporates measured water chemistry (e.g., pH, Eh, and
major ion chemistry) throughout the plume to assess the
potential for solubility control by a Pu(IV) or Am(lll) precipi-
tate such as a hydrous oxide phase. Identification of active
sequestration to prevent plutonium/americium migration in
ground-water provides justification for proceeding to Tier II
characterization efforts.
Tier II. Under Tier II, the apparent rate and mechanism(s)
of attenuation are determined. Estimates of a site attenu-
ation rate(s) can be assessed via a well transect along
the ground-water flow path. In addition, time-series data
may be collected at one or more monitoring points within
the plume (USEPA, 2007; Section IIIA.5). This informa-
tion will allow assessment of the relative timescales for
contaminant immobilization and fluid transport and deter-
mination of whether remediation objectives can be met
within the required regulatory time frame. In addition, the
mechanism(s) for attenuation are to be identified under
this stage of site characterization. This effort will require
determination of the chemical speciation of solid phase
plutonium/americium and may be approached according
to the following scheme:
1.) Determination of solution and solid phase
plutonium/americium concentrations, along with the
relative concentration of major ions/components in
aquifer solids where attenuation is occurring;
2.) Calculation of saturation state of ground water
relative to measured aqueous chemistry for
potential Pu(IV) or Am(lll) precipitates; and
3.) Determination of aquifer mineralogy (Amonette,
2002) to determine the relative abundance of
components that might support adsorption and/or
coprecipitation of plutonium/americium.
This compilation of information will facilitate identification of
the reaction(s) leading to plutonium/americium immobiliza-
tion. It is recommended that identification of redox-sensitive
components in aqueous and solid matrices be conducted
using samples collected in a manner that preserves their
in-situ speciation (USEPA, 2006b). The demonstration
of concurrence between conceptual and mathematical
models describing plutonium/americium transport will
entail development of site-specific parameterization of the
chemical processes controlling plutonium/americium solid
phase partitioning.
Tier III. Once the partitioning mechanism(s) have been
identified for the site, the subsequent characterization effort
under Tier III will involve determination of the stability of
immobilized plutonium/americium and the capacity of the
aquifer to sustain continued uptake. It is recommended
that the stability of immobilized plutonium/americium be
tested based on the anticipated evolution of ground-water
chemistry concurrent with decay of the plume. For example,
changes in ground-water pH and/or alkalinity can exert
a significant influence on the adsorption of plutonium/
americium onto aquifer solids. Therefore, it is recommended
that sediment leach tests be conducted to characterize
the magnitude of plutonium/americium mobilization as a
function of pH and alkalinity for a ground-water chemistry
representative of site conditions. It is recommended that
the capacity for plutonium/americium uptake onto aquifer
solids be determined relative to the specific mechanism(s)
118
-------
identified in Tier II. For example, if site characterization
under Tier II indicated that plutonium/americium adsorption
onto hydrous iron/manganese oxides was the predominant
attenuation process, then the mass distribution of these
mineral components along relevant ground-water flow paths
needs to be determined. This site-specific capacity would
then be compared to plutonium/americium mass loading
within the plume in order to assess the longevity of the natu-
ral attenuation process. If site-specific tests demonstrate
adequate stability of immobilized plutonium/americium and
sufficient capacity within the aquifer to sustain plutonium/
americium attenuation, then the site characterization effort
can progress to Tier IV. For cases where contaminant stabil-
ity is sufficient but aquifer capacity is insufficient for capture
of the entire plume, then a determination of the benefits of
contaminant source reduction is required.
Tier IV. Finally, under Tier IV a monitoring plan is estab-
lished along with contingency plans in the event of MNA
failure. It is recommended that the monitoring plan be
designed to establish both continued plume stability and to
identify changes in ground-water chemistry that may lead
to re-mobilization of attenuated plutonium/americium. The
specific chemical parameters to be monitored will include
those identified under Tier III that may halt plutonium/ameri-
cium partitioning and/or result in dissolution of either dis-
crete plutonium/americium precipitates or aquifer minerals
that sequester plutonium/americium from ground water. For
example, solution phase parameters that could alter either
plutonium/americium precipitation include inorganic carbon
(alkalinity), pH, and dissolved organic carbon concentration.
Changes in these parameters may occur prior to observed
changes in solution plutonium/americium and, thus, serve
as monitoring triggers for potential MNA failure. In addition,
sites at which residual plutonium/americium sources are
left in unsaturated zones should include monitoring points
to assess changes in the release of these contaminants
to the saturated aquifer due to increased surface infiltra-
tion or rises in the ground-water table. Changes in system
hydraulics may serve as monitoring triggers for potential
MNA failure. In this instance, a contingency plan can
be implemented that incorporates alternative strategies
to arrest possible plume expansion beyond compliance
boundaries. Possible strategies to prevent plume expansion
include pump and treat operations, isolation or stabilization
of near-surface contaminant source zones, or installation of
reactive barriers to enhance uptake capacity perpendicular
to the direction of plume advance.
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122
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Appendix A - Radioactive Decay Processes
Basic concepts describing radionuclides and radioactive
decay can be obtained from several sources (USEPA,
2000; Siegel and Bryan, 2003). The term radionuclide
used throughout this document refers to an unstable atom
that releases energy and/or particles during conversion to
a new atom. An atom is composed of neutrons, protons,
and electrons, where electrons orbit around a nucleus
composed of neutrons and protons. Atoms with the same
number of protons, but different numbers of neutrons are
referred to as isotopes. For example, the element radium
has six isotopes; radium-223, radium-224, radium-225,
radium-226, radium-227 and radium-228. While each of
the radium isotopes contains 88 protons in their nucleus,
the number of neutrons for each isotope is different. The
variation in the number of neutrons does not change the
element or its chemical properties, but it can affect the sta-
bility of the element. This instability can result in radioactive
decay, leading to use of the term radioisotopes for unstable
isotopes. For consistency in this document, the term
radionuclide will be used when referring to radioisotopes.
Radionuclide X possesses an atomic number Z and mass
number A, and is represented as ZAX or AX where A = Z + N
and is the sum of the numbers of protons (Z) and neutrons
(N) in the nucleus (e.g., radium-223 is alternatively repre-
sented by 223Ra). Radioactivity is the property of spontane-
ous emission of particles or electromagnetic radiation from
an unstable nuclide. The principal modes of radioactive
decay include transformation via emission of alpha, beta,
and positron (positively charged beta particle) particles,
and via orbital electron capture. Following decay, if the
daughter nucleus is left in an excited state, the excess
energy may be shed through the emission of a gamma
ray or, alternatively, through the ejection of an inner shell
orbital electron - a process termed internal conversion. The
three primary types of radiation emitted during or following
radioactive decay are: alpha (a) and beta (p) particles and
gamma (y) rays. Examples of radioactive decay charac-
teristics for radionuclides addressed within this volume are
shown in Table 1.2.
Specific activity, the rate of radioactive decay per unit
mass, is fixed for any specific radionuclide regardless of
its chemical or physical state. However, the rate differs
greatly for different radionuclides. The decay rate is typically
expressed in terms of a half-life, which is the time required
for the radioactivity of a radionuclide to decay to one-half
of its original value. Half-lives for the different radionu-
clides vary from fractions of a second to billions of years.
Although chemical transformations are typically sensitive to
temperature, pressure, physical states, and other factors,
radioactive decay transformations are not.
Modes of Radioactive Decay
Radioactive decay occurs because the balance of neutrons
and protons in the nucleus of the nuclide is unstable. The
nuclide tries to reach a more stable configuration by emit-
ting particles and/or electromagnetic radiation (y-rays).
Nuclides emitting a particles (42He) decrease in A by four
units and in Z by two units:
238-rj
92 U
a-particles are usually emitted with between 3 and 9 MeV
of kinetic energy, but, since they are relatively massive
and doubly charged, they do not penetrate very far into
matter. A thick sheet of paper is sufficient to completely
stop a particles.
Beta (p) decay processes include electron emission (p~),
positron emission (p+) and electron capture (EC). The first
of these reactions, known as negatron decay or, simply,
as negative p-decay, can be represented as «->/>++e~ ,
where n = neutron, p+= proton, and e~ = electron. A specific
example of this decay process is the decay of 137Cs:
(electron emission or beta-minus decay)
The emitted electron is called the p particle. The result is
an increase in Z by one unit with no change in A. Unlike
the discrete energies of a decay, there is a broad, continu-
ous distribution of energies, extending from almost zero to
some maximum value from p decay. If negative p-particle
emission occurs for nuclei possessing a high (and unstable)
N/Z value, then a reverse process might be expected to
occur - specifically decay should convert a proton to a neu-
tron. As in negatron decay, this process may be repeated
in several consecutive steps before a stable N/Z value is
obtained. Conversion of a proton to a neutron can occur
in two different ways - either by emission of a positron
(positive beta decay, where the reaction is /r -» n + e+;
e+ = positron) or by absorption of an electron, usually from
the K or L shells of the atom (electron capture, where the
reaction is p+ + e -» n). Examples of these two decay
processes include:
C — >
+ (positron emission or beta-plus decay) and
2°
Am + e~ -> 2g°Pu (electron capture)
123
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Fission decay can be either spontaneous or neutron-
induced. Spontaneous fission is a naturally occurring
process in which a nucleus breaks into two fragments,
along with the emission of 2-3 neutrons. An example of a
spontaneous fission decay process is:
29fCf
9
2 Nd+2 neutrons
The neutron-induced fission process is initiated by captur-
ing a neutron, as in:
235
92
Ba + Kr + 2 neutrons
The nuclides formed in a fission process are called fission
products. Each fissionable nuclide can produce a wide
range of fission products, many of which are not radioac-
tive. In a group of identical radionuclides, several decay
modes may be observed. The competition is expressed
by the branching ratios which correspond to the relative
probability of occurrence of a decay mode. In general, the
branching ratio for a particular decay mode is defined as
the ratio of the number of atoms decaying by that decay
mode to the number decay in total. As an example, plu-
tonium-241 (241Pu) decays to americium-241 (241Am) and
uranium-237 (237U) via beta emission and alpha emission,
respectively. However, the probability for decay to241 Am is
much higher and thus dominates the distribution of decay
products (i.e., the fraction of 237U produced during 241Pu
decay is approximately 2.45x10'5; Appendix A in USEPA,
1993b, EPA/402/R93/081). Discussion of the significance
of branching to decay for radionuclides addressed in this
volume is provided in the individual contaminant chapters.
Modes of Nuclear De-excitation
(Following Decay)
Often, the daughter nucleus resulting from radioactive
decay is left in an excited state. The nucleus can shed this
excess energy, or de-excite, via two processes: gamma ray
emission and internal conversion (1C). The first process,
gamma emission, involves the removal of excess energy
with the emission of a gamma ray. Gamma rays are elec-
tromagnetic radiation similar to X rays, ultraviolet and visible
light, and radio waves. Emitted y-rays decrease the mass
of the nucleus by an amount corresponding to the energy
carried away by the y-rays. For example, in a decay of
238U, 77% of the a particles have 4.18 MeV and 23% have
4.13 MeV of energy. Decay by emission of 4.13 MeV a
particles leaves the nucleus with 0.05 MeV greater energy
than do 4.18-MeV a emissions. This 0.05 MeV difference
is accounted for through emission of a y ray of that energy.
Emission of y-rays occurs immediately (< 10~12s) follow-
ing a- or p-decay, but in some cases the nucleus may
remain in the higher energy state for a measurable length
of time (milliseconds or greater). The excited state of the
nucleus and its daughter state are referred to as a nuclear
isomer. Gamma rays are emitted with a discrete energy,
like a-particles, rather than with a continuous spectrum of
energies as negatron and positron decay particles are in
p-decay. Since they have no mass or charge, y-rays do
not interact readily with matter and therefore exhibit greater
penetration in air and matter than charged particles. The
second process, internal conversion, involves energy trans-
fer from the nucleus to an inner electron shell, resulting in
the ejection of a high-energy electron. As with decay by
orbital electron capture, the inner shell electron vacancy
is quickly filled by an outer shell electron and also results
in emission of an X-ray.
It should be noted that a number of radionuclides decay
without leaving an excited daughter nucleus, and therefore
do not result in gamma emission or internal conversion.
These isotopes are known as pure emitters. Several beta
emitters, including 3H, 14C, 32P, 90Sr, and 90Y, decay without
gamma emission. The absence or presence of gamma
emission is an important consideration for radiation detec-
tion and measurement and for radiation protection.
Decay Chains
Most naturally occurring radioactive materials and many
fission products undergo radioactive decay through a series
of transformations rather than in a single step. Until the last
step, these radionuclides emit energy or a particle(s) with
each transformation and become another radionuclide.
Man-made elements, which are all heavier than uranium
and unstable, undergo decay in this way. This decay chain,
or decay series, ends in a stable nuclide. For example,
the decay chain for uranium-238 (238U) is illustrated in
Figure A.1. Uranium-238 has the longest half-life, 4.5 billion
years, and polonium-214 the shortest, 164 microseconds.
The last radionuclide in the chain, polonium-210 (210Po)
transforms to the stable nuclide, lead-206 (206Pb). The fol-
lowing labeling scheme is used in Figure A.1 and throughout
this document to illustrate radioactive decay processes:
Stable isotopes exist for some of the elements addressed
in this volume (e.g., strontium; 84Sr, 86Sr, 87Sr, 88Sr). As
the name implies, stable isotopes do not undergo radio-
active decay. It is important to note that a stable isotope
for a given element may not be derived from decay of an
unstable isotope of the same element. As an example, the
chapter on strontium, included in this volume, focuses on
discussion of 90Sr as a radionuclide commonly encountered
in groundwater plumes at sites with radioactive wastes
(USEPA, 1993a). Radioactive decay of 90Sr leads to the
production of 90Y (beta decay), which subsequently decays
to stable 90Zr. Thus, stable isotopic versions of strontium
present within a plume may be derived from a source
other than that releasing 90Sr into the subsurface. Stable
isotopes of strontium that may be present within a plume
are likely derived from natural sources within the aquifer
(e.g., due to weathering of mineral components within aqui-
fer solids). From a site characterization perspective, it is
important to understand that unstable and stable isotopes
for a given element may occur within a plume, since this
may govern the types of contaminant detection methods
that are employed to define transport of the radionuclide
targeted for remediation.
124
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Isotope
(mass
number &
element)
Decay
by alpha
particle
emission
Decay
half-life
& units
Decay
by
electron
emission
\
^^v
***••
a
* 222Rn
3.82 d
•^ a
218Po
^ 3.05m
a
214pb
26.8 m
(0.9998)*
218At |
2.00s
(0.0002) !
K" *t ^
P-
a
+
/
214Bi
19.9m
2
P-^X
/r
P-
^
Z+1
z
Z-1
r
\
Branching
fractions
^
N+1
N-1
N-2
x-ax\s corresponds with
change in number of neutrons
(decreasing to the right)
y-sx\s corresponds with
change in atomic number
(decreasing down)
arrows point in "direction" of
decay process
solid lines correspond to the
predominant decay reaction
dashed lines correspond to a
branching decay reaction of
secondary importance
There are several sources for tabulated data characterizing
decay of radionuclides. The source used by the USEPA in
risk calculations is derived from the ICRP38 Nuclear Decay
Data Files (e.g., Appendix A in USEPA, 1993b, EPA/402/
R93/081). Another source of decay data is the Evaluated
Nuclear Structure Data File (ENSDF; http://www.nndc.bnl.
gov/ensdf/}, which contains evaluated nuclear structure
and decay information for over 2900 nuclides. Technical
evaluations of these data are published in Nuclear Data
Sheets (http://www.nndc.bnl.gov/ndsA. which is a journal
primarily devoted to the publication of evaluated nuclear
structure and decay data.
Units and Specific Activity
The Curie (Ci) has been used as a radioactive decay unit
for many years and is still being used widely in the United
States, but the unit used under the International System
of Units (SI), Becquerel (Bq), has practically replaced the
Curie unit and is the only unit accepted by most scientific
publishers. The two are related as follows:
1 Becquerel (Bq) = 1 disintegration s'1
1 Curie (Ci) = 3.7x 1010 s'1 (Bq) or
1 Bq = 27 pCi (p = 10'12)
The specific radioactivity describes the relationship between
radioactivity and mass and is the decay rate (counts per
unit of time) per unit mass of a substance. It is necessary
to make clear whether the mass refers to a pure radionu-
clide or to a mixture - the SI unit is Bq kg-1. For practical
purposes specific radioactivity is also defined in dpm g-1
or dpm mole'1. Activity concentration (or "radioactive con-
centration") is given in Bq rrr3 or Bq M. For example, the
total atoms in 1 g of 32P (t1/2 = 14.3 days) is:
1
1/32 moles of 32P
(6.023 x 1023)/32 atoms of 32P.
From the decay equation:
dN/dt = -AN
where N is the number of radioactive atoms, t is time (s)
and A, is the decay constant (0.693/t1/2). The activity can
be obtained from:
dN/dt =
0.693 x6.022x!023
' 32x14.3x24x60x60
(-)1.055 x 1016 disintegrations
disintegrations s
where the minus sign indicates that the total number of
atoms in the sample, N is decreasing overtime. Therefore,
1 g of pure 32P has the activity of 1.055 x 1016 Bq (or
2.853 x 105Ci).
With a half-life of 1599+4 y, the specific activity per gram
of 226Ra is 0.988 Ci or 3.7 x 1010 Bq or 2.9 x 1012 dpm.
The specific activities of some of the longer-lived naturally
occurring radioactive species are: 40K, 31.3 kBq kg-1; 232Th,
4.05 MBq kg-1; and 238U, 12.4 MBq kg-1.
125
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238 |J
4.47x109y
a
r
234Jh
24.10d
234pa
6.70 h
*
f3"
234U
2.44x105y
P"
1
a
r
230Jh
7.70x104y
a
226Ra
1.60x103y
a
222Rn
3.82d
3.050 m
238 U
Decay
Series
a
26.8 m
(0.9998)
2.00s
(0.0002)
'- - t - •
P~ '
"a
214Bj
19.9m
a
22.30y
210Bj
5.01d
210
138 d
a
206
stable
Figure A.1 Decay chain for238U showing intermediate nuclides formed during series transformation to stable 208Pb. Half-
lives for radioactive decay were obtained using the WinChain program that provides electronic access to the
ICRP38 Nuclear Decay Data Files (ICRP, 1983; Eckerman et ai, 1994); \is = microseconds, s = seconds,
m = minutes, d = days, y=years. Decay types are shown as a for alpha particle emission and p- for electron
emission. Decay branching is shown for218Po with the branching fraction for both daughter products (214Pb
and218At) shown in parentheses (a decay dominates).
126
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In general, the longer-lived radionuclides decay slowly and
are more persistent in the environment. However, their
specific activity is relatively low. Shorter-lived radionuclides
decay more quickly to stable forms, and their specific activity
is generally much higher than those of longer-lived radio-
nuclides. The daughter products of a number of decaying
radionuclides are themselves radionuclides, which can
also provide significant doses of radiation. For example,
90Sr decays to 90Y which subsequently decays to stable
90Zr. The contribution from 90Y is approximately 8% of that
from 90Sr based on drinking water contaminated with 90Sr
and 90Y in equilibrium.
References
Eckerman, K.F., R.J. Westfall, J.C. Ryman, and M. Cristy.
Availability of nuclear decay data in electronic form,
including beta spectra not previously published. Health
Physics 67:338-345 (1994).
ICRP (International Commission on Radiological Protection).
Radionuclide Transformations: Energy and Intensity of
Emissions. ICRP Publication 38. Annals of the ICRP
16:2-3. International Commission on Radiological
Protection, Pergamon Press, New York (1983).
Siegel, M.D. and C.R. Bryan. Environmental Geochemistry
of Radioactive Contamination. SAND2003-2063,
Sandia National Laboratories, Albuquerque, NM (2003).
http://www.prod. scmdia. gov/cgi-bin/techlib/access-
control.pl/2003/032063.pdf
USEPA. Environmental Characteristics of EPA, NRC, and
DOE Sites Contaminated with Radioactive Substances,
EPA 402-R-93-011, Office of Radiation and Indoor Air,
Washington DC (1993a).
USEPA. External Exposure to Radionuclides in Air, Water,
and Soil, Federal Guidance Report No. 12, EPA402-R-
93-081, Office of Radiation and Indoor Air, Washington
DC(1993b).
USEPA. Radionuclides Notice of Data Availability Technical
Support Document, Office of Ground Water and
Drinking Water, Washington DC (2000). http://www.
epa.gov/safew ater/radionuclides/pdfs/regulation_
radionuclides rulemakins techsupportdoc.pdf
127
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