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

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
ii

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.
iii

Foreword

The U.S. Environmental Protection Agency is charged by Congress with protecting the Nation’s land, air, and water 

resources. Under a mandate of national environmental laws, the Agency strives to formulate and implement actions 

leading to a compatible balance between human activities and the ability of natural systems to support and nurture 

life. To meet this mandate, EPA’s research program is providing data and technical support for solving environmental problems today and building a science knowledge base necessary to manage our ecological resources wisely, 

understand how pollutants affect our health, and prevent or reduce environmental risks in the future.

The National Risk Management Research Laboratory (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. NRMRL’s research 

provides solutions to environmental problems by: developing and promoting technologies that protect and improve 

the environment; advancing scientific and engineering information to support regulatory and policy decisions; and 

providing the technical support and information transfer to ensure implementation of environmental regulations and 

strategies at the national, state, and community levels.

This publication has been produced as part of the Laboratory’s strategic long-term research plan. It is published and 

made available by EPA’s Office of Research and Development 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

v

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
vi

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
vii

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
viii

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
ix

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 ICRP. ........................................................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 aSO4

2- 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 SO4

2-. ..................................95

Figure 9.5 Solubility of ThO2(am) over variable log PCO2 values (-4 to -2) and 1 mM SO4

2-. ....95
x

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

Figure 10.1 Decay reactions involving 135Cs and 137Cs. .............................................................103

Figure 11.1 Decay series for 239Pu. ............................................................................................112

Figure 11.2 Decay series for 240Pu. ............................................................................................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 = 10-2.5 atm. .............................115

Figure 11.5 Solubility diagram for Am(III) at PCO2 = 10-2.5 atm and PO2 = 0.2 atm. .................115

Figure A.1 Decay chain for 238U showing intermediate nuclides formed during series 

transformation to stable 206Pb. .................................................................................126
xi

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 Affiliation

Robert G. Ford USEPA/ORD, National Risk Management Research Laboratory, Cincinnati, OH 

45268

Richard T. Wilkin USEPA/ORD, National Risk Management Research Laboratory, Ada, OK 74820

Daniel I. Kaplan Savannah River National Laboratory, Aiken, SC 29808

James E. Amonette Pacific Northwest National Laboratory, Fundamental Science Directorate, Richland, 

WA 99352

Patrick V. Brady Sandia National Laboratories, Geochemistry Department (MS-0750), Albuquerque, 

New Mexico 87185

Paul M. Bertsch University of Kentucky, Lexington, KY 40506

Kenneth Lovelace USEPA/OSWER/OSRTI, Washington, DC 20460 (deceased)

Stuart Walker USEPA/OSWER/OSRTI, Washington, DC 20460

Ronald Wilhelm USEPA/OAR/ORIA, Washington, DC 20460

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

Craig Bethke University of Illinois, Department of Geology, Urbana, IL 61801

Douglas B. Kent 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).

xiii

Executive Summary

The term “monitored natural attenuation,” as used in this document and in the Office of Solid Waste and Emergency 

Response (OSWER) Directive 9200.4-17P, refers to “the reliance on natural attenuation processes (within the context 

of a carefully controlled and monitored site cleanup approach) to achieve site-specific remediation objectives within 

a time frame that is reasonable compared to that offered by other more active methods.” When properly employed, 

monitored natural attenuation (MNA) may provide an effective knowledge-based remedy where a thorough engineering analysis informs the understanding, monitoring, predicting, and documenting of the natural processes. In 

order to properly employ this remedy, the Environmental Protection Agency needs a strong scientific basis supported by appropriate research and site-specific monitoring implemented in accordance with the Agency's Quality 

System. The purpose of this series of documents, collectively titled “Monitored Natural Attenuation of Inorganic 

Contaminants in Ground Water,” is to provide a technical resource for remedial site managers to define and assess 

the potential for use of site-specific natural processes to play a role in the design of an overall remedial 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 nonradionuclide 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 characterization 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 situations 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). 

1

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 contaminant 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, radon, technetium, thorium, tritium, strontium, and uranium. 

The selection of these contaminants was based on two criteria. First, a selected element had to be one of high priority 

to the site remediation or risk assessment activities of the 

USEPA (USEPA, 1993a; USEPA, 2002a; USEPA, 2006a; 

USEPA, 2007c). Second, selection was based on chemical 

behavior considering chemical traits such as: toxicity, cations, anions, conservatively transported, non-conservatively 

transported, and redox sensitive elements (USEPA, 1999b; 

USEPA, 2004a). Using these characteristics of the contaminants, 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 radionuclide contaminants for this document is representative 

of these characteristics. 

1.1.2 Purpose of Document

This document is intended to provide a technical resource 

for determining whether MNA is likely to be an effective 

remedial approach for inorganic contaminants1 in ground 

water. This document is intended to be used during the 

remedial investigation and feasibility study phases of a 

Superfund cleanup, or during the equivalent phases of a 

RCRA Corrective Action (facility investigation and corrective 

measures study, respectively). The decision to select MNA 

as the remedy (or part of the remedy) will be made in a 

Superfund Record of Decision (ROD) or a RCRA Statement 

of Basis (or RCRA permit).

The USEPA expects that users of this document will include 

USEPA and State cleanup programs and their contractors, 

especially those individuals responsible for evaluating alternative 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 contaminants are discussed in this document. There are two reasons 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 evaluating 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 contaminated 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 generic 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 remedial technologies at the site. “Conceptual site model” is not synonymous 

with “computer model”; however, a computer model may be helpful for 

understanding and visualizing current site conditions or for predictive 

simulations of potential future conditions.

Conceptual Background for Radionuclides 
2

capacity of the aquifer to sustain attenuation of the contaminant 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 intended to address policy considerations or specific regulatory 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 example, 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 overseeing 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 reference 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 ‘natural 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 destruction of contaminants. (USEPA, 1999c, page 3.)

Even though several physical, chemical, and biological 

processes are included in the above definition, the1999 

OSWER Directive goes on to state a preference for those 

processes that permanently degrade or destroy contaminants, 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 unacceptable to the overseeing regulatory authority. 

Therefore, sites where the contaminant plumes 

are no longer increasing in extent, or are shrinking, would be the most appropriate candidates 

for MNA remedies.” (USEPA, 1999c, page 18.)

Control of contaminant sources is also an important aspect 

of EPA’s policy. The actual policy language is given below:

“Control of source materials is the most effective means of ensuring the timely attainment of 

remediation objectives. EPA, therefore, expects 

that source control measures will be evaluated 

for all contaminated sites and that source control measures will be taken at most sites where 

practicable. At many sites it will be appropriate 

to implement source control measures during the 

initial stages of site remediation (“phased remedial approach”), while collecting additional data 

to 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:
3

MNA may, under certain conditions (e.g., through 

sorption or oxidation-reduction reactions), effectively reduce the dissolved concentrations and/or 

toxic forms of inorganic contaminants in groundwater 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 uranium) 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 concentration, 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, radionuclides 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 

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.

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 processes. Often, however, they may exist in forms that 

have low mobility, toxicity, or bioavailability such 

that they pose a relatively low level of risk. Therefore, 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 applicable or relevant and appropriate requirements (ARARs) 

unless a waiver is justified. Cleanup levels for response actions 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 standards include federal maximum contaminant levels 

(MCLs) and/or non-zero maximum contaminant 

level goals (MCLGs) established under the Safe 

Drinking Water Act (SDWA), or more stringent state 

drinking water standards. Other regulations may 

also be ARARs as provided in CERCLA §121(d)

(2)(B).

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 

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. Unlike 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.
4

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 (μg/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 

30 pCi/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.gov/superfund/health/contaminants/radiation/nuclides.htm.

Radionuclide

Contaminant

Current MCLa or UMTRCA 

(pCi/L)

Mass Equiv to MCL, UMTRCA, or 

RBL (μg/L)

Americium 241Am 15 0.0000044

Cesium 137Cs 200 0.0000023

Tritium 3H 20,000 0.0000021

Iodine 129I 1 0.0057

Plutonium

238Pu 15 0.00000088

239Pu 15 0.00024

240Pu 15 0.000066

241Pu (27 RBL)b 0.00000026

242Pu 15 0.0038

244Pu 15 0.85

Radium

226Ra 5c 0.0000051

228Ra 5c 0.000000018

Strontium 90Sr 8 0.000000059

Technetium 99Tc 900 0.053

Thorium

228Th 15 0.000000018

229Th 15 0.000071

230Th 15 0.00074

232Th 15 140

Uranium

234U 30d 30d

238U 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 emitters; MCL is 15 pCi/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 10-6 risk. These were calculated using equation 11 in Risk Assessment Guidance 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 pCi/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 mg/L for uranium; UMTRCA ground-water standard 

is 30 pCi/Lcombined for U-234 and U-238.
5

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 contaminated 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 radioactive material during radioactive decay and absorbed by 

exposed biological tissue is the key driver of health effects 

from exposure to most radionuclides. In addition, radioactive 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 technologies 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) = 2.8 x 10-15 * A * T1/2 * MCL(pCi/L)

where 2.8 x 10-15 for water is a conversion factor, A is the 

radionuclide atomic weight in g/mole, and T1/2 is the radionuclide 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 characterize the quantities of radioactive material in the environment 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 environment 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 potassium. 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 represent 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 
6

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 specific 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 ‘reduce 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 reactions 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 precipitation, 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 immobilization 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.

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.
7

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 impeded 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 relative to the velocity of ground-water flow (e.g., 222Rn (radon), 

3.82 day half-life). Where the rates of decay and fluid transport 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., 99Tc 

(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 characteristics 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 disparity in radioactive decay rate for the radionuclides that are 

encountered at contaminated sites. Further discussion of 

these issues is provided later in this document.

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) shrinking 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/activity exceeds Risk-based or ARAR criterion. The symbol “ t” refers to the characteristic time for transport 

(ttrans) and radioactive decay (tdecay).
8

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. (A) Time to achieve ground-water cleanup standards 

following an order-of-magnitude activity reduction for radionuclides addressed in this volume. (B) Time 

series plots of decay-only activity reductions to achieve cleanup standard in ground water for an orderof-magnitude change in total activity for several radionuclides addressed in this volume. (C) Influence of 

total initial activity on the time to achieve representative cleanup standard (example is for 90Sr in one liter 

of solution); blue hatched area shows region where a relevant mass equivalent cleanup standard may be 

achieved within 100 years. 

For several radionuclides discussed in this volume, reactions with dissolved or solid phase components within the 

aquifer lead to transfer of the contaminant from ground 

water onto aquifer solids. In this situation, characteristics 

of the contaminant plume may be governed by immobilization processes. In essence, there are two effective ‘plumes’ 

within the plume boundary: 1) the dissolved and mobile 

plume (including the mobile colloidal phase), and 2) the 

solid phase and immobile plume (Figure 1.4). As with 

non-radionuclides, the overall stability of the immobilized 

contaminant to resist re-mobilization becomes a critical 

factor for the viability of natural attenuation as a cleanup 

remedy. The stability of the immobilized radionuclide will 

be governed by the inherent chemical characteristics of the 

contaminant, the chemical characteristics of ground water, 

and the stability of aquifer solids to which the contaminant 

may be bound. Changes in ground-water chemistry over 

time may result in conditions that cause contaminant release or dissolution of host aquifer solids. Evaluation of this 

issue may be further complicated for radionuclides whose 

decay products possess inherent chemical properties that 

differ significantly from that of the parent radionuclide.
9

‘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 product(s)

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 ( ttrans > tdecay ), 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 addition, 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 designing laboratory tests and reaction transport models used 

to assess and project potential for re-mobilization of the 

radionuclide. An overview explaining the types of immobilization 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 contaminant 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 significant 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 prioritizes and limits the data that is needed for decision making 

at each screening step. Conceptually a tiered analysis approach seeks to progressively reduce uncertainty as sitespecific data are collected. The decision-making approach 
10

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 collection and evaluation within the tiered analysis approach 

would be structured as follows:

I. Demonstration that the ground-water plume is not 

expanding and that sorption of the contaminant onto 

aquifer solids is occurring where 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. 

1.3.1 Tier I

The objective under Tier I analysis would be to eliminate 

sites where site characterization indicates that the groundwater 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 incorporation 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 objective would be to demonstrate contaminant uptake onto 

aquifer solids. Analysis of ground-water plume behavior 

at this stage is predicated on adequate aerial and vertical 

delineation of the plume boundaries. Characterization of 

ground-water plume expansion could then be supported 

through analysis of current and historical data collected 

from monitoring wells installed along the path of groundwater 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 scenarios 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 activity 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 collection and analysis of contaminant concentrations on the 

retrieved aquifer solids. Illustration of the type of data trend 

anticipated for a site where sorption actively attenuates 

contaminant transport 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.

1.3.2 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 remediation”.) Data collection and analysis performed for Tier II 

would indicate whether MNA processes are capable of 

achieving remediation objectives, based on current geochemical 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 recommended 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, 
11

in each hydrostratigraphic unit. The necessary data might 

include physical parameters such as hydraulic conductivities 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, aquifer solids composition and mineralogy, and the chemical 

speciation of the contaminant in ground water and associated 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 governing contaminant attenuation. Indirect microbial influence 

on contaminant attenuation includes situations in which the 

predominant characteristics of the ground-water chemistry 

are controlled by microbial oxidation-reduction reactions. 

This situation may be more predominant in plumes in 

which readily degradable organic contaminants, such as 

hydrocarbons or chlorinated solvents, are also present. 

Ultimately, mechanistic knowledge of the attenuation process 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.

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 pCi/L and (B) 1-month 

release from source at 50,000 pCi/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.
12

1.3.3 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 objectives 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 contaminants, 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 typically sensitive to this parameter. Alternatively, dissolution of 

a contaminant attenuated as a carbonate precipitate may 

result from decreases in ground-water pH and alkalinity.

Assessment of attenuation capacity will depend on knowledge of the flux of contaminants and associated reactants 

in ground-water, as well as the mass distribution of reactive 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 distribution 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 example, 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 conditions 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 laboratory tests employing aquifer solids collected from within the 

plume boundaries that can be exposed to solutions that 

mimic anticipated ground-water chemistries (e.g., ambient ground-water samples or synthetic solutions in which 

the concentrations of specific dissolved constituents can 

be systematically varied). A supplementary avenue to test 

contaminant stability could include use of chemical reaction 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 boundaries supports natural attenuation processes with sufficient 

efficiency, capacity, and stability. The technical knowledge 

obtained through identification of the specific attenuation 

mechanism and the sensitivity of the attenuation process to 

changes in ground-water chemistry can then be employed 

in designing a monitoring program that tracks continued 

performance of the MNA remedy.

1.3.4 Tier IV

The objective under Tier IV analysis is to develop a monitoring program to assess long-term performance of the 

MNA remedy and identify alternative remedies that could 

be implemented for situations where changes in site conditions could lead to remedy failure. Site data collected during 

characterization of the attenuation process will serve to 

focus identification of alternative remedies that best match 

site-specific conditions. The monitoring program will consist 

of establishing a network of wells: 1) that provide adequate 

aerial and vertical coverage to verify that the ground-water 

plume remains static or shrinks, and 2) that provide the 

ability to monitor ground-water chemistry throughout the 
13

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 network 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 attenuation 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 or sulfide precipitates, respectively. Non-contaminant 

performance parameters such as these will likely serve as 

“triggers” to alert site managers to potential remedy failure 

or performance losses, since the attenuation reaction will 

respond to these changed conditions. Since increases in 

mobile contaminant concentrations may be delayed relative 

to changes in site conditions, these monitoring parameters 

may improve the ability of site managers to evaluate and 

address the potential for ground-water plume expansion. In 

summary, the tiered analysis process provides a means to 

organize the data collection effort in a cost-effective manner that allows the ability to eliminate sites at 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 hydrology, ground-water chemistry, contaminant distribution, and 

the aqueous speciation of the contaminant. These system 

characteristics can have direct influence on the selection of 

pump-and-treat or in-situ remedies best suited to achieve 

cleanup 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 selection. The remaining discussion in this section of Volume 1 

will elaborate on two issues that have been introduced 

above, specifically the use of models in site characterization and general factors to consider for implementation of 

a long-term performance monitoring program. These topics are addressed at this juncture to allow greater focus to 

discussions later in this volume pertaining specifically to 

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 contaminant 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 applications, 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 significantly 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 example, 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 undesirable situation if the daughter radionuclide presents a similar 
14

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

I

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

II 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)

III

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
15

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 parent has much shorter half-life than unstable daughter (e.g., 

241Am 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 characteristics different from that of the parent radionuclide.

Figure 1.6 Illustration of four decay in-growth scenarios that may be encountered for ground-water plumes contaminated with radionuclides. Illustrations were derived from the USEPA website - http://www.epa.gov/radiation/understand/equilibrium.html. T1/2,p = decay half-life of parent radionulide, T1/2,d = decay half-life of 

daughter radionuclide.
16

1.4.1.2 Data Sources and Determination of SolidLiquid Partitioning

As described in Section IIIB.3 in Volume 1 (USEPA, 2007a), 

there are several different approaches to describing solidliquid partitioning in transport models. The simplest approach to modeling solid-liquid partitioning during groundwater 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 developed 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 conditions 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 partitioning (e.g., surface complexation models; see Section IIB in 

Volume 1). However, the applicability of these models is 

usually limited to describe radionuclide transport in contaminant 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 characteristics of sorbent phase(s) along transport pathways 

is addressed via collection of site-specific data, while allowing 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 thermochemical 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 groundwater systems. Examples of these efforts include those by 

the Nuclear Energy Agency (Guillaumont et al., 2003) and 

the Paul Scherrer Institute Laboratory for Waste Management (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 radionuclide 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 results 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 
17

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; Maher et al., 2006). However, this phenomenon is 

not routinely recognized as a contributing factor in groundwater plumes at contaminated sites. 

 In general, the impact of this process is difficult to characterize 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 alpharecoil events should be qualitatively evaluated before ruling 

out the significance of this process. In other instances, 

consideration of alpha-recoil processes may be more relevant from the standpoint of explaining anomalous analytical results and/or selecting/tailoring an analytical method.

1.4.2 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 colloidal solids acting as carriers of inorganic contaminants 

within a ground-water plume. The types of colloids discussed 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 technical 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 mobile 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 appropriate 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 enhanced contaminant transport in place of direct observation 

of colloids along plausible transport pathways. In general, it 

appears that the common approach of identifying the presence 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 composition 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 evidence 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 expansion 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 technologies 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, selection of MNA depends on developing detailed understanding 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 characterization of site hydrology (Section IIIA), 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 IIIB.2). These characterization tasks 

also apply to the radionuclides discussed within this volume 

and specific recommendations are included in the individual 
18

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 detection 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 provides 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 radiation) possessing energies that are also specific to the decay 

process (e.g., as reviewed in USEPA, 2006b). Thus, detection 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 concentration 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 particle 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 application 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 methodologies (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/reportSAM030107.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/radiation/marlap/manual.html. Analysis of liquid samples (or solids following dissolution) using the radiometric techniques 

described above requires significant sample processing 

1) to isolate radionuclides of interest from potentially interfering 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 environmental 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 (Quantification of Radionuclides; USEPA, 2004b). This document also 

provides discussion of the use of liquid scintillation methods in which decay of a radionuclide excites a compound 

(“fluor”) that produces fluorescent radiation.

It is recommended that site managers also consult a concise 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 document, 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).
19

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 characteristics for detection and quantification of radionuclides in 

ground water. In situ sensors for long-term monitoring applications without use of consumable reagents have also 

been developed (e.g., Egorov et al., 2001) for the quantification of 99Tc 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., 234/235/238U, 239/240/244Pu, 99Tc, 129I), 

mass-based techniques may be faster and more sensitive 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 135/137Cs. 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 determination of radionuclide concentrations in environmental 

samples. The sensitivity and mass-selectivity of these approaches, 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 identify 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 radionuclide continuously decreases in time due to radioactive 

decay (in the absence of production from decay of progenitors 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 corrections 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 radionuclides 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 concentration 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.

1.5.2 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)], although the uranyl cation may adsorb 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 oxidation 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 speciation 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 aqueous speciation of long-lived radionuclides in environmental 

samples, including evaluation of the analytical merits of the 

various methods for determining contaminant speciation.

1.5.3 Multiple Sources for Radionuclide of 

Concern

Characterization of radionuclide transport within a plume 

is commonly evaluated through the installation and sampling 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 concentration of the radionuclide in space and time relative 
20

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 gradient 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 Section 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 disposal 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 evaluated 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 radionuclides 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 technetium 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 differentiating 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 leaching reactions driven by movement of the plume through 

the aquifer. For example, isotopic ratios of 235U:238U or 236U:238U in 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 processing 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 

236U:238U atom ratio of 10-14 (Zhao et al., 1997). The ratio of 

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 235U:238U is 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 235U:238U is 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 
21

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 plutonium 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 ingrowth of 239Pu or 240Pu are illustrated in Table 1.3. The relative 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 relative 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 ingrowth. 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 effort. 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.bnl.gov/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, b ̄ = 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

Decay

Progenitor

Decay

Mode

Progenitor

Decay Half-life

239Pu

239Np b ̄ 2.355 d

243Cm a 28.5 d

243Am (via 239Np) a (b ̄) 7380 y

240Pu

240Np b ̄ 65 m

244Cm a 18.11 y

244Am (via 244Cm) b ̄ (a) 10.1 h

1.5.4 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 present 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 installation 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 activity. 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 development of the well (e.g., surging and/or high pumping rates) 

can be used to remove these types of solids from within 
22

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 materials 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 redeveloped, 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 Hassellöv 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 finegrained iron oxyhydroxides in ground water with elevated 

ferrous iron concentrations [>10 mg/L Fe(II)] is an indication 

that iron oxidation-precipitation reactions may be occurring 

during sample collection or processing. The use of procedures 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; 

Hassellöv 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 mm 

membrane filter. Thus, use of a conventional filtration pore 

size such as 0.45 mm 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 progressed 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 approaches that reliably sample colloidal forms of contaminants from ground water. Ultimately, it is recommended 

that ground-water sampling for the purpose of evaluating 

the presence of mobile colloids be conducted using permanent monitoring points that have a clear record of well 

construction (including description of granular solid materials employed during drilling and screen placement within 

the formation), procedures used for screen development, 

and the data used to evaluate the adequacy of well development and sampling procedures.

1.5.5 Aquifer Solids

The objectives and methodologies presented in 

Sections IIIB and IIIC 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 realtime 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).
23

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26
27

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 1×10-15 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 movement 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 evaluations 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 radioactive 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 pathways for tritium.

Attenuation 

Processes

Mobilization 

Processes

Characterization 

Approach

Radioactive 

decay

Not applicable 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 interferences. 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 
28

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, measurement 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 characterization 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 pathways 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 groundwater cleanup.

Tier I. Site characterization under Tier I will involve demonstration that the ground-water plume is static or shrinking, 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, determination 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 information, 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 attenuation (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 infiltration 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 contaminant source areas. Changes in land usage (including 

at recharge zones, wetlands, and over the contaminated 
29

site) and/or ground-water withdrawals from the aquifer may 

influence ground-water flow direction and velocity (USEPA, 

2007; Section IIIA), which in turn directly influences 3H travel 

time. It is recommended that additional tritium transport 

modeling be included to evaluate the impact of these various 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 established 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 strategies to arrest possible plume expansion beyond compliance 

boundaries. Possible strategies to prevent plume expansion include the installation of barriers to minimize tritium 

migration from source areas and/or ground-water extraction 

with surface treatment.

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Research 43:W10434, doi:10.1029/2006WR005847 

(2007).

Visser, A., J.D. Schaap, H.P. Broers, and M.F.P. 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, P.E., 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).
31

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 alpharecoil 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 naturallyoccurring 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 uranium 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 shortlived 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 accumulation 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 redoxcontrolled 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 subsurface 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 radioactive 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 concentrations. 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 
32

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 (b ̄), or electron capture (e). 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.

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 (b ̄). 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.
33

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 pathways for radon.

Attenuation 

Processes

Mobilization 

Processes

Characterization 

Approach

Radioactive 

decay

Not applicable 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 groundwater sampling and processing methods that result in collection 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 radonin-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 dissolved 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 reasonably 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 daughter 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 ascertain 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 dissolved 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 contaminant 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 
34

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 influence 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 groundwater cleanup.

Tier I. Site characterization under Tier I will involve demonstration that the ground-water plume is static or shrinking, 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, determination 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 information, in combination with knowledge of contaminant source 

location(s), radon mass, and emanation characteristics 

from source zones, can be employed to develop a decaytransport model to project radon activity/concentration 

distribution throughout the plume. The demonstration of 

concurrence between conceptual and mathematical models 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 capacity 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 groundwater 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 IIIA), which in turn directly influences radon travel 

time. It is recommended that additional radon transport 

modeling be included to evaluate the impact of these various 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 established 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 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 the removal of 226Ra from 

periodically saturated source areas and/or ground-water 

extraction with surface treatment.
35

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37

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 reprocessing. Otherwise, the primary inputs of 90Sr to the environment 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 relatively 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(II)]. 

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 fission 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 mg/L (Jacobson 

and Wasserburg, 2005) and 14-43,500 mg/L (McIntosh 

and Walter, 2006). This compares to the mass-equivalent 

risk based concentration limit for 90Sr of 0.000000059 mg/L 

(see Table 1.1).

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 (b ̄). 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 cation 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 synthetic 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 strontium will likely be limited to near-surface systems with high 

concentrations of dissolved organic carbon.
38

Solubility

Strontium may precipitate with dissolved carbonate or sulfate 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-bearing 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 minerals in combination with surface complexation reactions on 

oxyhydroxide minerals (e.g., Jackson and Inch, 1989).

Figure 4.2 (a) Solubility and speciation of strontium 

as a function of pH and PCO2 (strontianite, SrCO3). (b) Strontianite stability field 

(Sr 2+=10 -6m, 0.087 mg L-1) in relation to stability fields for calcite (CaCO3 , variable Ca 2+

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. 
39

Site Characterization

Overview

Attenuation of 90Sr might be achieved through radioactive 

decay with the influence of adsorption or coprecipitation reactions. In general, since strontium adsorption is 

dominated by highly 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. 

Co-precipitation reactions are likely to occur only under situations 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 pathways for strontium.

Attenuation 

Processes

Mobilization 

Processes

Characterization 

Approach

Radioactive 

decay

Not 

applicable

Determination 

of ground-water 

velocity along 

relevant transport 

pathways and 

contaminant mass 

release rate from 

source areas.

Adsorption 

onto aquifer 

minerals (clay 

minerals, iron 

oxyhydroxides)

Desorption 

due to 

decreasing 

pH or 

competition 

from major 

cations 

in ground 

water.

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.

Co-precipitation 

with Fe(II) 

carbonate 

or calcium 

carbonate

Dissolution 

due to 

decreasing 

pH or 

oxidation 

of Fe(II) 

carbonate

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 quantified using radiometric analysis (USEPA, 2004; USEPA, 

2006). However, the presence of other beta-emitting 

radionuclides in the sample can interfere with the determination of the activity/concentration of 90Sr via detection of 

beta emissions due to overlap in their corresponding betaemission 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 chromatographic 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 online concentration of 90Sr in the analyzed sample (similar to 

methods employed for radiometric techniques) and through 

use of dynamic reaction cell configurations to control formation 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 introduced 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 
40

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 reactions. 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 retardation reactions limit strontium migration. Bulk solid-phase 

partitioning can be conducted using total digestion or acidextractable 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 betaemitters, could be significant due to release of these constituents 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 exchangeable 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 fractionation or selective extraction of carbonate minerals, 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 IIIB).

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 displace 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 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 strontium sequestration onto 

aquifer solids and the long-term stability of solid phase 

strontium as a function of existing and anticipated groundwater 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 demonstration 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 collected 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 distributed 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 
41

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 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). 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 adsorption 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) Calculation of saturation state of ground water 

relative to precipitation of carbonate or (hydr)oxide 

minerals along relevant 90Sr transport pathways;

2) 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 demonstration 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 determined, 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 present 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 immobilization, 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 established 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 90Sr and 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 concentrations 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 perpendicular 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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45

Technetium

Patrick V. Brady, James E. Amonette, Robert G. Ford, Richard T. Wilkin

Occurrence and Distribution

99Tc 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 99Tc, 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, naturally 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 decommissioning 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 drinking water, one can estimate an equivalent activity-based 

MCL of 900 pCi/L for 99Tc (mass equivalent concentration 

of 0.053 mg/L) assuming 99Tc as the only beta- or photonemitting radioisotope (USEPA, 2002). 

Aqueous Speciation

99Tc can exist in eight oxidation states ranging from Tc(-I) 

to Tc(VII). The two most common oxidation states are 

Tc(VII) and Tc(IV) under oxidizing and reducing conditions, 

respectively. Reviews of the aqueous geochemistry of technetium 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 alkalinity (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 5.1a 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 5.1b). However, many of the mixed- or lower-oxidation 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(IV)-carbonate 

complexes could limit precipitation of this hydrous oxide 

in systems with alkaline pH and high carbonate alkalinity. 
46

Available thermodynamic databases list solubility constants 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(-II)2(S(-I)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 observations 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(II) 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(IV) 

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 typically does not adsorb onto aquifer solids. The sorption Kd 

for 99Tc 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) to Tc(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 apparent retardation within an aquifer is due to the latter process.

Figure 5.1 Phase stability diagrams for technetium at 

25°C. a) Eh-pH diagram for the system Tc-H2O, 

Tc = 10-7.5 (53,903 pCi/L), showing metastable 

field for TcO2· 2H2O. b) Eh-pH diagram for 

the system Tc-H2O, Tc = 10-10 to 10-8 (170-

17,045 pCi/L), all phases unsuppressed (LLNL 

database). 
47

Figure 5.2 Phase stability diagrams for technetium at 25°C. 

a) Eh-pH diagram for the system Tc-H2O-CO2 , 

Tc = 10-7.5 (53,903 pCi/L), showing metastable 

field for TcO2· 2H2O. Tc(IV)-carbonate complex 

data from Hummel et al. (2002). b) Eh-pH 

diagram for the system Tc-H2O-S; Tc = 10-7.5 

and S = 10-3.

Site Characterization

Overview

Attenuation of 99Tc might be achieved through precipitation 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 99Tc 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 99Tc concentrations 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, 99Tc will be present as the 

soluble TcO4 ̄ anion, which is only slightly adsorbed to the 

solid phase. As a beta particle emitter (Ebmax = 294 keV), 99Tc 

can be quantified using scintillation counting techniques 

(USEPA, 2004b; USEPA, 2006a). Interference from other 

beta-emitting radionuclides, e.g., 90Sr (Ebmax = 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 99Tc 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 pertechnetate 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 99Tc 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 (Ebmax = 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 measure 99Tc 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 99Tc 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 sensitive and accurate detection of 99Tc in complex aqueous 
48

matrices (e.g., Richter et al., 1997; Eroglu et al., 1998; Kim 

et al., 2002). The potential for carryover of 99Tc 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 99Tc attenuation 

within the aquifer and the capacity along relevant transport 

pathways include the determination of 99Tc partitioning to 

aquifer solids and identification of aquifer solids mineralogy that may participate in reduction or precipitation reactions. Evaluation of the mass distribution of 99Tc between 

co-located ground water and aquifer solids throughout the 

plume provides an assessment of the extent that retardation 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; USEPA, 2006a). Because the 

overall quantities of solid-phase 99Tc are likely to be low, 

techniques for determination of total 99Tc or the chemical 

fractionation of 99Tc in aquifer solids is needed. Fusion 

methods that convert the solid matrix to a form easily dissolved 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 99Tc 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 99Tc speciation in sediments (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 information on analysis approaches and analytical techniques 

applied to solid phase characterization is provided in USEPA 

(2007; Section IIIB). There are no methods that have been 

developed to identify the oxidation state of extracted 99Tc 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 99Tc at concentrations 

much higher than typically encountered in ground water. 

The performance of analytical procedures used to quantify 

released 99Tc 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 99Tc 

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 attenuated 99Tc and the capacity for the aquifer to sustain 99Tc 

removal from ground water include the persistence of 

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(II)-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.
49

reducing conditions and the mass flux of 99Tc 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 99Tc 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 situation where the flux of 99Tc 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 99Tc 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 

99Tc, radioactive decay will not provide a viable mechanism 

for plume attenuation. Therefore, the goal of site assessment will be to demonstrate the process(es) controlling 

99Tc sequestration onto aquifer solids and the long-term 

stability of solid phase 99Tc as a function of existing and 

anticipated ground-water chemistry. 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 demonstration 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 collected to demonstrate 99Tc 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 99Tc precipitate such as a hydrous 

oxide phase. Identification of active sequestration to prevent 99Tc 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 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). 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 solid phase 99Tc 

and may be approached according to the following scheme:

1) Determination of solution and solid phase 99Tc 

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(II) associated with aquifer solids) 

and/or coprecipitation; and 

4) Determination of 99Tc-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 99Tc immobilization. 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 

99Tc transport will entail development of site-specific parameterization of the chemical processes controlling 99Tc 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 99Tc and the capacity of the aquifer to sustain 

continued uptake. It is recommended that the stability of 

immobilized 99Tc 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 the solubility of a hydrous 

Tc(IV) oxide. Therefore, it is recommended that sediment 

leach tests be conducted to characterize the magnitude 

of 99Tc mobilization as a function of pH for a ground-water 
50

chemistry representative of site conditions. It is recommended that the capacity for 99Tc 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 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 99Tc mass 

loading within the plume in order to assess the longevity of 

the natural attenuation process. If site-specific tests demonstrate the stability of immobilized 99Tc and that there is 

sufficient capacity within the aquifer to sustain 99Tc 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 established 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 99Tc. The specific 

chemical parameters to be monitored will include those 

identified under Tier III that may halt 99Tc partitioning and/or 

result in dissolution of either discrete 99Tc precipitates or 

aquifer minerals that sequester 99Tc from ground water. 

For example, solution phase parameters that could alter 

either 99Tc 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 99Tc 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 components that induce more reducing conditions. 

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Vaughan, R.A.D. Pattrick, W.A. Reed, and L.N. Moyes. 

X-ray absorption spectroscopy studies of reactions of 

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coupled plasma mass spectrometry. Analytica Chimica 

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in aquifer sediments. Geomicrobiology Journal

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a radiometric preconcentrating minicolumn sensor in 

an equilibration-based sensing approach. Analytical 

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preparation of technetium 99m sulfide colloid for 

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(2008).
53

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 uranium deposits are comprised of a few common minerals 

including oxides (uraninite and pitchblende), silicates (coffinite, 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 nonnuclear 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 μg U/L, but range as high as 

2000 μg U/L near natural uranium deposits in arid regions. 

In streams, levels of uranium are typically about 1 μg U/L, 

but values as high as 50 μg U/L are seen in granitic watersheds. The level of uranium in seawater is 3 μg 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) processing 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 subsurface 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 systems (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.
54

Table 6.1 Illustration of potential decay paths from different progenitor sources leading to production of uranium 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, b ̄ = 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 Decay Progenitor Decay Mode Progenitor Decay Half-life

238U 242Pu a 3.763 x 105 y

235U

235Np EC 396.1 d

239Pu a 24065 y

234U

234Np EC 4.4 d

234Pa b- 6.70 h

238Pu a 87.74 y

238Np → 238Pu b ̄ → a 2.117 d → 84.74 y

236U

236Np 1 EC 153000 y

240Pu a 6537 y

1 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). 
55

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 (b ̄). The half-life is shown directly below the isotope that 

is subject to decay; y = years, d = days, h = hours, m = minutes, s = seconds, μs = 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 of 218Po follows two routes resulting in 98.98% production 

of 214Pb and 0.02% production of 218At).
56

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 predominant 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(III), are rare and generally 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 thermochemical 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 et al., 2002; GormanLewis 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 carbonato complexes of the uranyl (UO2

2+) 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 influence 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 μg 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 groundwater systems, although field evidence suggests that this 

would be significant only in more acidic (pH<6) systems 

(e.g., Ranville et al., 2007).

Figure 6.2 Distribution of aqueous U(VI) species among oxyhydroxyl-, sulfate-, carbonate-, and calcium carbonatecomplexes as a function of ground-water chemistry. (a) Dependence on alkalinity expressed as mol/L 

CaCO3(aq) ; pH = 7, 0.005 mol/L NaCl, 0.001 mol/L K2SO4 , 0.001 mol/L MgNO3, 0.42 μmol/L U (100 μg U/L). 

(b) Dependence on pH; 0.001 mol/L CaCO3(aq) (100 mg/L CaCO3 alkalinity), 0.005 mol/L NaCl, 0.001 mol/L 

K2SO4 , 0.001 mol/L MgNO3 , 0.42 μmol/L U (100 μg U/L). Model predictions using Visual MINTEQ 

Version 2.53 (Based on MINTEQA2 described in Allison et al. (1991); available at http://www.lwr.kth.se/

English/OurSoftware/vminteq/). Precipitation of Ca/Mg carbonates suppressed; slight oversaturation for 

pH > 8 and/or > 0.005 M CaCO3(aq).
57

Solubility

Under ambient conditions, the thermodynamically stable 

uranium solid phases will be either U(VI) or U(IV) compounds. 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 insitu 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) concentrations 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; 

Arai 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 colloidalsized phases are not formed and transported. The most 

common uranium ore-forming process involves reductive 

precipitation of U(IV) phases as a result of microbiological 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=10-4.0. 

Bold lines show the pH-dependent solubility 

of soddyite [(UO2 )2SiO4·2H2O] at PCO2 =10-3.0

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 solubility of schoepite [ b-UO3·2H2O] at PCO2=10-3

bar. (b) Addition of Ca (10-4 to 10-2) and 

Na (10-3). Thermodynamic data for uranophane (Ca(H3O)2(UO2 )2(SiO4 )2·3H2O) are from 

Langmuir (1997; p. 552). Note that solubility 

studies by Pérez et al. (2000) suggest that 

uranophane may be less stable than depicted 

here.
58

Figure 6.4 Solubility of U(VI) as a function of pH and in 

the presence of phosphate (10-3.5), 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(UO2 )2 (PO4 )2 ), (UO2 )3 (PO4 )2 ·4H2 O(c), 

Schoepite, and UO2 HPO4 ·4H2 O(c). Dashed 

lines show the metastable extensions of aqueous U(VI) species. Thermodynamic data for 

U-P complexes and solids are from Guillaumont 

et al. (2003). 

Figure 6.5 Eh-pH diagrams for U at 25°C. (a) System 

U-O-H-Si-C, with aH4 SiO4 = 0-4.0 and 

PCO2 = 10-2.5 bar. (b) System U-O-H-C 

with SU = 10-5 and PCO2 = 10-2.5 bar. 

(c) System U-O-H-C-Na-Ca with SU = 10-5, 

aH4 SiO4 = 10-4.0, and PCO2 = 10-2.5 bar. 

Soddyite = (UO2 )2 SiO4 ·2H2 O; urani- O; uraninite = UO2 ; schoepite = b-UO3 ·2H2 O; uranophane = Ca(H3 O)2 (UO2 )2 (SiO4 )2 ·3H2 O.
59

Adsorption

Adsorption of uranium typically involves inner-sphere complexation of uranyl (i.e., those containing UO2

2+) 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 geochemistry 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 reaction with aquifer solids and/or external recharge/infiltration 

from atmospheric precipitation or surface water. As previously 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 partitioning (e.g., Sato et al., 1997; Um et al., 2007). It has been 

demonstrated that changes in ground-water chemistry influence 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 intermediate 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 species 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 contaminated 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 sediments 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 concentrations in environmental samples. The sensitivity and 

mass-selectivity of these approaches, 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., 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 electrochemical sensor that facilitates selective extraction 

and detection of the uranyl anion [U(VI)] from natural 

water samples with a 5 mg/L detection limit. The sensor 

is constructed using an ion implanted polymer that can 
60

be incorporated into optical detection configurations for 

screening or quantification of uranyl concentrations in aqueous 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 UO2

2+-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 uraniumcontaminated soils. Alternatively, detection of the uranyl 

dioxo-cation (UO2

2+) can also be achieved in the laboratory 

using methods such as laser-induced kinetic phosphorimetry (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 detection 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 (234U:238U, 235U:238U, or 236U:238U) 

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 

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.
61

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 (Arai 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 absorption 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 determination 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 determination of specific radioisotope abundance in the liquid 

extracts has been used to identify potential sources of 

uranium contamination. Howe et al. (2002) used extraction 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 determine the influence of discharge from a nuclear fuel enrichment plant. Total uranium concentrations in sediments 

decreased with distance from the suspected source, and 

the 238U/235U ratio 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 solutions 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 conducted under a nitrogen atmosphere. These authors also 

demonstrated that this solution did not extract U(IV) from 

reduced sediment samples under nitrogen, so this procedure 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 contaminated 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 bicarbonate 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., phosphates 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 uranium 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 
62

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 remobilization 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 groundwater 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 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 demonstrate 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 groundwater 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 demonstration 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 collected 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 characterization efforts. 

Tier II. Under Tier II, the apparent rate and mechanism(s) 

of attenuation are determined. Estimates of a site attenuation rate(s) can be 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 speciation 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 precipitation of ferrous iron in anoxic ground water). The demonstration of concurrence between conceptual and mathematical 

models describing uranium transport will entail development 
63

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 sustain 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 degradation 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 compared 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 capacity 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 established 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 precipitates 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 components (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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69

Iodine

Daniel I. Kaplan, Robert G. Ford, Richard T. Wilkin 

Occurrence and Distribution

Iodine occurs as a trace element in the Earth’s crust (average 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 radioisotopes 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 129I/127I ratio from natural sources is substantially lower 

than that from human nuclear activity (Hou et al., 2009). 

Common anthropogenic sources of 129I include the production 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 subsurface 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 1.57x107 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 I- 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), I2(aq) (oxidation state 0; 

“iodine”) can be produced from the oxidation of I ̄ or the 

reduction of IO3 ̄. Iodine partitions between a soluble species [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 mercury (Gammons and Yu, 1997; Pruszyn ́ski 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 mediated 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(II)-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(II) 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 

I

2(aq) 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 
70

may be preceded by the reduction of IO3 ̄ (e.g., Steinberg 

et al., 2008a). Iodide (I ̄) may also be oxidized to IO3 ̄ during 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 

et al., 2009).

Figure 7.1. Eh-pH diagram of dominant iodine aqueous 

species at 25°C (total I = 10-6 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 emissions 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 predominant 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 important 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 relative 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 experiments 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 presence 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 
71

active throughout the plume. A list of potential attenuation 

processes is provided in Table 7.1. 

For some sites, the volatile transfer of I2(g) or methylated 

species of 129I from shallow ground water into overlying 

unsaturated soil or directly into open or confined atmospheres may also reduce ground-water concentrations. 

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 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 pathways for iodine.

Attenuation 

Processes

Mobilization 

Processes

Characterization 

Approach

Adsorption of 

IO3 ̄ or I ̄ onto 

aquifer minerals 

(clay minerals, 

iron/manganese 

oxyhydroxides)

Desorption 

due to 

increasing pH 

or competition 

from major 

anions in 

ground water.

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).

Covalent 

bonding to 

immobile forms 

of NOM in 

aquifer solids

Degradation or 

mineralization 

of NOM 

resulting in 

release of 

complexed 

iodine; 

dissolution/

desorption 

of NOM into 

ground water

Determine concentration 

of aqueous and solid 

forms of NOM and 

bound 129I along 

transport pathway(s).

Aqueous Measurements

Since 129I decays by emitting b-particle with a maximum 

energy of 154.4 keV and g-rays of 39.6 keV as well as X-rays 

(29–30 keV), it can be measured using g-spectrometry 

and b-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 spectrometry (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 interference from formation of molecular ions within the plasma 

such as 127I

1H1H+ 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 129I/127I mass concentration 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 particulates in the sample (e.g., Buraglio et al., 2000). It is recommended 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 species 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 speciation 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 anionexchange 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.
72

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 contamination 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 et al., 2000b). 

Distinguishing between stable (127I) and contaminant forms 

of iodine (129I) partitioned to aquifer solids may be necessary 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, including evaluation of individual isotopes in reference soil/sediment 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 combustion 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 documented. For example, an updated measurement for the 

concentration of 129I in the IAEA-375 reference soil material 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 129I/127I 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 extraction of iodine from the sediment by heating in an oxygen 

atmosphere. However, as discussed by Hou and Roos 

(2008), reliable determination of I129/I127 ratios <10x-10 (i.e., 

a pre-nuclear age ratio) may require the use of accelerator 

mass spectrometry. Analysis of the 129I/127I ratio, in addition 

to the total 129I concentration, provides a means to demonstrate 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 concentration, 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 process, 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 Um 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 desorption 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 groundwater 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 
73

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-ofmagnitude, 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 significant 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 μg/L and 129I = 0.45 μg/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 hydrogeologic 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 sorption 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 groundwater cleanup. 

Tier I. Site characterization under Tier I will involve demonstration 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 collected 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 geochemical 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 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). 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 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 129I-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 demonstration of concurrence between conceptual and mathematical 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 evolution 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 function 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 
74

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 various 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 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 recommended.

Tier IV. Finally, under Tier IV a monitoring plan is established 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 99Tc 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 oxyhydroxides). 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 parameters 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 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 

components that induce more oxidizing conditions.

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79

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 controlled 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 

~0.8 x 10-6 mg/kg. Radium concentrations in igneous rocks 

increase by 100 times along the geological sequence: ultrabasic (0.009 x 10-6 mg/kg), basic igneous (0.6 x 10-6mg/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 containing 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 products 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 et al., 1995; Rutherford et al., 1996; Kim et al., 

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), uranium 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 luminescent 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 relative abundance of 224Ra and 228Ra may be increased, relative 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 mechanism 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).
80

Figure 8.1 Production of 228Ra, 224Ra, and 223Ra (and the radon daughters of 224Ra and 223Ra) from the decay of 236U and 235U that may 

be present in ground water at levels greater 

than natural abundance due to uranium enrichment activities. Decay modes include those 

leading to ejection of an alpha particle (a) or a 

beta particle (b-). 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 of 227Ac 

follows two routes resulting in 98.6% production 

of 227Th and 0.014% production of 223Fr). 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 

O a k R i d g e N a t i o n a l L a b o r a t o r y 

(http://ordose.ornl.gov/downloads.html).

Figure 8.2 Decay series for 226Ra based 

on data from ICRP (1983). Decay modes include 

those leading to ejection of an alpha particle 

(a) or a beta particle (b-). The half-life is shown 

directly below the isotope that is subject to 

decay; y = years, d = days, m = minutes, 

s = seconds, ms = 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 of 218Po follows two routes 

resulting in 98.98% production of 214Pb and 

0.02% production of 218At). 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).
81

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 thermodynamic data of aqueous and solid speciation of radium 

was reviewed by Langmuir and Riese (1985). Ra2+ has little 

tendency to form aqueous complexes and of the alkalineearth 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 Guo, 

1998; Sturchio et al., 2001; USEPA 2004). As illustrated 

by the analyses presented in evaluations of radium aqueous 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 sulfuric 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, RaCl2, 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 Guo, 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 typically 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 Guo, 1998; Martin and Akber, 1999; 

Grundl and Cape, 2006).

Figure 8.3 pH – log aSO4

2- 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 10-3 m to 10-1.8 m. 

b) PCO2=10-3 bars and variable Ba2+ from 10-7 m to 10-5 m. 

At the MCL for Ra (2.2e-13 m), the Ra carbonate and Ra 

sulfate end-members are highly undersaturated at PCO2=10-3 

bars and log a SO4

2- from -6 to -1; however, Ra could be 

sequestered by Ba and Ca sulfates and carbonates.
82

Adsorption

Recent overviews of radium adsorption processes in subsurface 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 patterns 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 suggested 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 formations 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 aquifer 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 coprecipitation or adsorption dependent on the prevailing 

ground-water chemistry within the plume and the relative 

abundance and stability of immobile sorbent phases associated with aquifer solids. While radioactive decay may 

be a viable process for the short-lived radioisotopes, i.e., 

223Ra, 224Ra, and 228Ra, this will likely not serve as a viable 

attenuation process for plumes dominated by the longerlived 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/carbonate 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 conducted 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 groundwater 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 

oxyhydroxides)

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.
83

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 radiumenriched 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 radioisotope 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 (1x10-9 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 quantification 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 groundwater 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 samples 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 techniques, using ICP-MS as the analytical platform, have 

become increasingly reliable and sensitive. Hou and Roos 

(2008) have reviewed issues of potential analytical interference 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 currently 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 groundwater 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, 226Ra 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 transport pathways include the determination of partitioning of 

radium radioisotopes to aquifer solids, the cation-exchange 

capacity of aquifer solids, and identification of aquifer solids mineralogy that may participate in adsorption and/or 

coprecipitation reactions. Evaluation of the mass distribution 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 fractions using either non-destructive (i.e., X-ray fluorescence 

for stable elements and alpha/gamma spectrometry for 
84

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 spectrometry 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 for 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 SaintFort et al. (2007), it may be possible to use field-screening 

detection methods, based on measurements of total radiation 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 ionexchange 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 concentration 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 partitioned to readily accessible ion exchange sites, weak acid 

extractable phases (e.g., calcite), and/or reducible iron/

manganese oxides (Pardue and Guo, 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 sediment 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 geochemical 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 assessing 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 information on analysis approaches and analytical techniques 

applied to solid phase characterization is provided in USEPA 

(2007; Section IIIB).

Long-Term Stability and Capacity 

The long-term stability of radium attenuated through coprecipitation 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 Guo, 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 

other terminal electron acceptors [e.g., solid phase Fe(III) or 

Mn(IV)] consumed as a result of stimulated microbial activity. 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 remobilization 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 
85

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 sequestration 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 radioisotopes. 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 demonstration 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 collected 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 distributed between ground water and aquifer solids throughout 

the plume is recommended to account for both existing 

and potentially mobile forms of radium. This field partitioning 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 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). 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 (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 coprecipitation 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 relative 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 carbonate 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 identification 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 chemical processes controlling radium solid phase partitioning.

Tier III. Once the contributions from radioactive decay 

and adsorption/coprecipitation processes have been determined, 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 
86

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 groundwater chemistry representative of site conditions. It is recommended 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 stability 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 established 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 concentrations and/or increased concentrations of competing cations 

in ground water. As an example, increases in barium concentrations in ground water could signal either 1) the potential 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(II), 

Mn(II), 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 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 perpendicular 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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91

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)PO4] 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 thoriumcontaining 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 commercial 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 μg/L or 0.1 pCi/L 232Th (Hem, 1985), and it 

is even lower in oceans, averaging ~5.3 x 10-5 μg/L (Huh 

and Bacon, 1985). As will be discussed below, these concentrations 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 concentrations in excess of 2000 μg/L were reported in acidic 

ground waters beneath uranium tailings sites (Langmuir 

and Herman, 1980). In another example, natural aqueous 

thorium concentrations of >22 μg/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 isotopes that may be of concern for waste disposal operations 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 radioisotope 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. 
92

Figure 9.1 Decay series for 229Th including parent 233U. The 

half-life is shown directly below the isotope that 

is subject to decay; y = years, d = days, h = hours, 

m = minutes, s = seconds, ms = microseconds, 

stable = non-radioactive isotope. Decay halflife 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 9.2 Decay series for 234Th and 230Th including 

parents 238U and 234U. Production of 226Ra and 

“radon” (222Rn) 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, ms = 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).
93

Figure 9.3 Decay series for 232Th including parent 236U 

and progeny 228Th. Production of 228Ra, 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, ms = 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).

Aqueous Speciation

Although other oxidation states of thorium can exist under 

extreme laboratory settings, only Th(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 species 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)2

2+, Th(OH)3+, 

and Th(OH)4

0(aq) 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 concentrations (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 increasing 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 relatively 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 carbonate has been reported to form strong complexes with 

thorium to form ternary species with the general species 

formula Th(OH)y(CO3) z

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 general, it has been proposed that formation of complexes with 

dissolved natural organic matter could dominate thorium 
94

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 wetland 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 mm). 

Characterization of the size distribution of DOC and thorium concentrations following further size-fractionation of 

<0.2 mm-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 potentially 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 solubility 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 dissolved 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 dissolved 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 formation 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., Guo 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/fulvic compounds 

will likely depend on the relative concentration of dissolved 

organic matter and available adsorption sites on aquifer 

solids, with adsorption inhibition (or desorption) becoming 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 groundwater systems.
95

Figure 9.4 Solubility of ThO2(am) in (a) pure water and 

(b) 1 mM SO4

2-. Diagrams constructed at 

25°C using the LLNL thermodynamic database (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(SO4)2·9H2O.

Figure 9.5. Solubility of ThO2(am) over variable log PCO2

values (-4 to -2) and 1 mM SO4

2-. Diagram 

constructed at 25°C using the LLNL thermodynamic database (therm.com.v8.R6+) modified 

using data reported in Altmaier et al. (2006) for 

Th-OH-CO3 complexes and ThO2 (am) solubility, 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 coprecipitation or adsorption dependent on the prevailing 

ground-water chemistry within the plume and the relative 

abundance and stability of immobile sorbent phases associated with aquifer solids. While radioactive decay may be a 

viable process for the short-lived radioisotopes, i.e., 228Th or 234Th, this will likely not serve as a viable attenuation process 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.
96

Table 9.1 Natural attenuation and mobilization pathways for thorium.

Attenuation 

Processes

Mobilization 

Processes

Characterization 

Approach

Radioactive 

decay for 

short-lived 

radioisotopes 

(e.g., 228Th, 234Th)

Not applicable

Determination of 

ground-water velocity 

along relevant 

transport pathways 

and contaminant mass 

release rate from source 

areas.

Precipitation

Dissolution 

due to 

changes in 

ground-water 

chemistry

Evaluate formation of 

hydrous thorium oxide 

along relevant transport 

pathways based on 

calculated chemical 

saturation state of ground 

water.

Adsorption 

onto aquifer 

minerals (clay 

minerals, iron/

manganese 

oxyhydroxides; 

immobile 

forms of 

natural 

organic 

matter)

Desorption 

due to 

decreasing 

pH, increases 

in alkalinity or 

complexation 

by organic 

ligands 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 thorium partitioned to 

sorbent phases in aquifer 

solids along relevant 

transport pathways.

Aqueous Measurements

Available methods for the determination of thorium radionuclides 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 spectrometry 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 radioisotopes 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 interfering 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 precipitation 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 measurements for ground-water samples, isolation of the dissolved 

thorium radioisotopes from the sample matrix is commonly 

employed to improve the accuracy of radioisotope determinations. 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 uraniumbearing 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 225U:235U 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 determination 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 extraction 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 reasonable surrogates. Spiking aquifer solids with these surrogate 

phases provides the basis for assessing 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 thorium solid-phase partitioning. Examples of 
97

procedures to identify the type and abundance of specific 

minerals along transport pathways are available in the literature (e.g., Amonette, 2002). Additional information on 

analysis approaches and analytical techniques applied to 

solid phase characterization is provided in USEPA (2007; 

Section IIIB).

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 overlying 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 facilitate 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 colloidfacilitated 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 physical 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 attenuation 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 demonstration 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 collected 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 thorium 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 characteristics 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 
98

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. If concentrations of shortlived thorium radioisotopes (e.g., 228Th, 234Th) are elevated 

within a plume, it may be important 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 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 adsorption onto existing aquifer minerals or newly formed minerals 

predominates. Measurements of changes in ground-water 

chemistry along a transport pathway, e.g., decreasing ferrous iron concentrations coupled to increasing dissolved 

oxygen concentrations, may be used to infer formation of 

newly formed minerals. Confirmation of thorium adsorption 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. 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;

2. 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 chemical 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 thorium 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 alkalinity 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 thorium re-mobilization as a function of pH for a ground-water 

chemistry representative of site conditions. It is recommended 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 established 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 chemical 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 precipitates. 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 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 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 

adsorption capacity within the aquifer through the injection 

of soluble reactive components that precipitate within the 

formation.
99

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103

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 production 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 present 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 (CsAlSi2O6; 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 = 2.3x106 years), and 137Cs (t1/2 = 30.0 

years) (Isnard et al., 2009). The stable isotope, 133Cs, predominates 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 

137mBa (95%). In 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. 

Figure 10.1 Decay reactions involving 135Cs and 137Cs. 

Radioactive decay for 137Cs occurs via both a 

metastable intermediate (137mBa) and directly 

to 137Ba via ejection of a beta particle (b-). The 

decay of 137mBa to 137Ba involves emission of a 

662 keV gamma (g) 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 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).

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 lowtemperature 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 structures and high cation exchange capacity, including 2:1 layersilicate 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 
104

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 reactions. 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 radioactive 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 pathways for cesium.

Attenuation 

Processes

Mobilization 

Processes

Characterization 

Approach

Radioactive 

decay

Not 

applicable

Determination of 

ground-water velocity 

along relevant 

transport pathways 

and contaminant 

mass release rate 

from source areas.

Adsorption or 

ion exchange 

onto aquifer 

minerals (clay 

minerals, 

layered 

manganese 

oxides)

Desorption 

due to 

decreasing 

pH or 

competition 

from major 

cations 

in ground 

water; 

reductive 

dissolution of 

manganese 

oxides.

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 b-particles with maximum energies of 514 keV (94.4%) and 1175 keV (5.4%), 

and it is accompanied by high-abundance (85.1%) g-ray 

emission with energy of 661.7 keV (USEPA, 2006; Hou and 

Roos, 2008). Due to the abundance and low self-absorption 

of the g-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 
105

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 aqueous 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 characteristics 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, dissolved 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 mm 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 IIIB).

The activity/concentration of 137Cs in shallow aquifers may 

be supported, in part, by contributions from historical atmospheric fallout from nuclear weapon testing and infiltration 

into the subsurface (e.g., Ciszewski et al., 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+240Pu 

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 adsorption (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 contaminated 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 contaminant 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 
106

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 demonstration 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 collected 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 distributed 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 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). 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 demonstration 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 determined, 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 relative 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 progress to Tier IV.

Tier IV. Finally, under Tier IV a monitoring plan is established 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 competing 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 chemistry 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 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 or installation of permeable reactive barriers to enhance uptake 
107

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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Environmental Protection Agency, Cincinnati, OH (2007). 

http://www.epa.gov/nrmrl/pubs/600R07139/600R07139.pdf

Valcke, E. and A. Cremers. Sorption-desorption dynamics 

of radiocaesium in organic matter soils. Science of The 

Total Environment 157: 275-283 (1994). 

Zachara, J.M., S.C. Smith, C. Liu, J.P. McKinley, R.J. Serne, 

and P.L. Gassman. Sorption of Cs+ to micaceous 

subsurface sediments from the Hanford Site, USA. 

Geochimica et Cosmochimica Acta 66:193-211 (2002).
110
111

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 processing 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 distributed throughout surface soils due to historical atmospheric 

testing of nuclear weapons, and the radioisotopes 240Pu, 241Pu, 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 majority 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 environment 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 to 241Am (t1/2 = 432 years) via emission 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 halflives 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, 

b- = 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

Decay

Progenitor

Decay

Mode

Progenitor

Decay 

Half-life

241Am

241Cm EC 32.8 d

241Pu b- 14.4 y

239Pu

239Np b- 2.355 d

243Cm a 28.5 y

243Am

(via 239Np) a (b-) 7380 y(2.355 d)

240Pu

240Am EC 50.8 h

240Np b- 65 m

244Cm a 18.11 y

241Pu 245Cm a 8500 y
112

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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).



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Figure 11.2 Decay series for 240Pu. The half-life is 

shown directly below the isotope that is subject to decay; y = years, h = hours, d = days, 

m = minutes, s = seconds, ms = microseconds, 

stable = non-radioactive isotope. Decay halflife 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).
113



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Figure 11.3 Decay series for 241Pu, including daughter 241Am. The half-life is shown directly below 

the isotope that is subject to decay; y = years, 

h = hours, m = minutes, ms = milliseconds, 

ms = 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).

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 oxidations 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(III)3+, 

Pu(IV)O2, Pu(V)O2

+, and Pu(VI)O2(CO3)22-. The distribution 

of plutonium species between reduced [Pu(III), 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 et al., 1998; Topin et al., 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 reactions 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 concentrations than plutonium. For shallow ground-water systems 

with elevated concentrations of natural dissolved organic 

carbon, plutonium may form soluble complexes with dissolved 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(III) 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 et al., 1995; Vercouter et al., 2005). As shown in 

Figure 11.5, carbonate complexes can dominate americium 

speciation in solution. Americium may also form complexes 
114

with humic/fulvic compounds (e.g., Kim et al., 1993). It has 

been demonstrated that formation of soluble complexes with 

humic/fulvic compounds can potentially facilitate transport 

of Am(III) in shallow ground water where surface recharge 

events can result in increased organic carbon concentrations and aquifer flow velocities (e.g., Marley et al., 1993; 

Artinger et al., 1998; McCarthy et al., 1998).

Solubility

Both Am(III) 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(III) or Pu(IV) near source 

zones where elevated concentrations of these radionuclides 

might be encountered. As shown in Figure 11.4, precipitation 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 indicate 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 thermochemical studies that predict hydroxycarbonates as the more 

stable phase for actinium(III) and lanthanum(III) metals 

(Merli et al., 1997). In addition, it has been shown that 

Am(III) readily substitutes for Ca(II) 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, formation 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 adsorption 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 solids. 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 species 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 governed 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 analytical measurements that may be conducted on sampled 

ground water and aquifer sediments to assist in identifying 

the attenuation mechanism(s) are discussed in the following paragraphs.

Aqueous Measurements

Overviews of radiometric techniques for determining the 

activity of plutonium and americium radioisotopes is provided 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 (2010a; 2010b). 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 
115

Figure 11.4 Eh-pH stability diagram for Pu at 25 °C and PCO2 = 10-2.5 atm. Diagrams are for solubility control by 

PuO2 and Pu(OH)4 (see text). Examples of the four oxidation states of plutonium include Pu(III)3+, Pu(IV)

O2, Pu(V)O2

+, and Pu(VI)O2(CO3)22-.

Figure 11.5 Solubility diagram for Am(III) at PCO2 = 10-2.5 atm and PO2 = 0.2 atm. 
116

Table 11.2 Natural attenuation and mobilization pathways for plutonium and americium.

Attenuation Processes Mobilizaton Processes Characterization Approach

Plutonium

Radioactive decay of 

241Pu 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(III).

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(III) 

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+240Pu. The sensitivity of these 

methods is generally good for these radioisotopes, but typically 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 

et al., 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 
117

Szechenyi (2008). Examples where plutonium radioisotope 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 characteristics 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 radiometric 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(III) 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(III/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 differences 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 contaminants 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 radionuclides 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 radioisotopes 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 plutonium 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 solubilization 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 available 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 concentrations 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/americium does not occur. For plutonium, a particular concern 

are situations in which plutonium is attenuated under 

reducing conditions that are induced by characteristics of 
118

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 plutonium/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 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. While 

radioactive decay will contribute to attenuation of 241Pu and 241Am, 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 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 demonstration 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 collected to demonstrate plutonium/americium 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 Pu(IV) or Am(III) precipitate 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 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). 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 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(III) 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 immobilization. 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) 
119

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 natural 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 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 established 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/americium partitioning and/or result in dissolution of either discrete 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 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 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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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 stability 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 Z

AX 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 represented by 223Ra). Radioactivity is the property of spontaneous 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 (b) particles and 

gamma (g) rays. Examples of radioactive decay characteristics 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 radionuclides 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 emitting particles and/or electromagnetic radiation (γ-rays). 

Nuclides emitting a particles (4

2He) decrease in A by four 

units and in Z by two units:

238 234 4

92 90 2 U Th He → +

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 (b) decay processes include electron emission (b-), 

positron emission (b+) and electron capture (EC). The first 

of these reactions, known as negatron decay or, simply, 

as negative b-decay, can be represented as n p e → ++ - , 

where n = neutron, p+ = proton, and e- = electron. A specific 

example of this decay process is the decay of 137Cs:

137 137

55Cs Ba 56 e → + -

(electron emission or beta-minus decay)

The emitted electron is called the b 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, continuous distribution of energies, extending from almost zero to 

some maximum value from b decay. If negative b-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 neutron. 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 p+ → 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:

( ) 11 11

6 5 C B positron emission or beta-plus decay e → + +

 and

( ) 240 240

95Am Pu electron capture 94 e- + →
124

Fission decay can be either spontaneous or neutroninduced. 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:

252 98 152

98 38 60 Cf Sr Nd → + + 2 neutrons

The neutron-induced fission process is initiated by capturing a neutron, as in:

235 1 140 94

92 56 36 U n 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 radioactive. 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, plutonium-241 (241Pu) decays to americium-241 (241Am) and 

uranium-237 (237U) via beta emission and alpha emission, 

respectively. However, the probability for decay to 241Am 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 (IC). The first process, 

gamma emission, involves the removal of excess energy 

with the emission of a gamma ray. Gamma rays are electromagnetic radiation similar to X rays, ultraviolet and visible 

light, and radio waves. Emitted g-rays decrease the mass 

of the nucleus by an amount corresponding to the energy 

carried away by the g-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 g ray of that energy. 

Emission of g-rays occurs immediately (≤ 10-12 s) following a- or b-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 

b-decay. Since they have no mass or charge, g-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 transfer 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 detection 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 following 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 radioactive 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 aquifer 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.
125

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/nds/), 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) = 3.7 × 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 radionuclide 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 concentration”) is given in Bq m-3 or Bq l-1. For example, the 

total atoms in 1 g of 32P (t1/2 = 14.3 days) is:

1 g = 1/32 moles of 32P

= (6.023 x 1023)/32 atoms of 32P.

From the decay equation:

dN/dt = -lN

where N is the number of radioactive atoms, t is time (s) 

and l is the decay constant (0.693/t1/2). The activity can 

be obtained from:

dN/dt = ( ) 23 0.693 6.022 10 1 disintegrations s 32 14.3 24 60 60

× × - - × × × ×

= (-)1.055 × 1016 disintegrations s-1

where the minus sign indicates that the total number of 

atoms in the sample, N is decreasing over time. Therefore, 

1 g of pure 32P has the activity of 1.055 × 1016 Bq (or 

2.853 × 105 Ci).

With a half-life of 1599±4 y, the specific activity per gram 

of 226Ra is 0.988 Ci or 3.7 × 1010 Bq or 2.9 × 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.

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238U

4.47x109 y 

234

U

2.44x10

5 y 

234

Pa

6.70 h 

234Th

24.10 d 

230

Th

7.70x104 y 

 

226

Ra

1.60x103 y 

 

222

Rn

3.82d 

218

At

2.00 s 

(0.0002)

 

218

Po

3.050 m 

214

Po

164 s 

210Po

138 d 

214

Bi

19.9 m 

210

Bi

5.01 d 

214

Pb

26.8 m 

(0.9998)

210

Pb

22.30 y 

206Pb

stable

238U

Decay

Series















  













238U

4.47x109 y 

234

U

2.44x10

5 y 

234

Pa

6.70 h 

234Th

24.10 d 

230

Th

7.70x104 y 

 

226

Ra

1.60x103 y 

 

222

Rn

3.82d 

218

At

2.00 s 

(0.0002)

 

218

Po

3.050 m 

214

Po

164 s 

210Po

138 d 

214

Bi

19.9 m 

210

Bi

5.01 d 

214

Pb

26.8 m 

(0.9998)

210

Pb

22.30 y 

206Pb

stable

238U

Decay

Series

238U

Decay

Series















  













Figure A.1 Decay chain for 238U showing intermediate nuclides formed during series transformation to stable 206Pb. Halflives for radioactive decay were obtained using the WinChain program that provides electronic access to the 

ICRP38 Nuclear Decay Data Files (ICRP, 1983; Eckerman et al., 1994); ms = microseconds, s = seconds, 

m = minutes, d = days, y = years. Decay types are shown as a for alpha particle emission and b- for electron 

emission. Decay branching is shown for 218Po with the branching fraction for both daughter products (214Pb 

and 218At) shown in parentheses (a decay dominates).
127

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 radionuclides. 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.sandia.gov/cgi-bin/techlib/accesscontrol.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, EPA 402-R93-081, 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/safewater/radionuclides/pdfs/regulation_

radionuclides_rulemaking_techsupportdoc.pdf

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