Monitored Natural Attenuation 

of Inorganic Contaminants in 

Ground Water

Volume 1 

Technical Basis for Assessment

EPA/600/R-07/139

October 2007

Monitored Natural Attenuation of Inorganic 

Contaminants in Ground Water

Volume 1 - Technical Basis for Assessment

Edited by

Robert G. Ford

Richard T. Wilkin

Robert W. Puls

U.S. Environmental Protection Agency 

Office of Research and Development 

National Risk Management Research Laboratory

Ada, Oklahoma 74820

Project Officer

Robert G. Ford

Ground Water and Ecosystems Restoration Division

National Risk Management Research Laboratory

Ada, Oklahoma 74820

National Risk Management Research Laboratory

Office of Research and Development

U.S. Environmental Protection Agency

Cincinnati, Ohio 45268
ii

Notice

The U.S. Environmental Protection Agency through its Office of Research 

and Development managed the research described here under EPA Contract 

No. 68-C-02-092 to Dynamac Corporation, Ada, Oklahoma, through funds provided 

by the U.S. Environmental Protection Agency’s Office of Air and Radiation and Office 

of Solid Waste and Emergency Response. It has been subjected to the Agency’s 

peer and administrative review and has been approved for publication as an EPA 

document. Mention of trade names or commercial products does not constitute 

endorsement or recommendation for use.

All research projects making conclusions or recommendations based on 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 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.

Stephen G. Schmelling, Director

 Ground Water and Ecosystems Restoration Division

National Risk Management Research Laboratory

v

Contents

Notice ............................................................................................................................................. ii

Foreword ........................................................................................................................................ iii

Figures .........................................................................................................................................viii

Tables ............................................................................................................................................ ix

Acknowledgments .......................................................................................................................... x

Executive Summary ...................................................................................................................... xi

Section I - Conceptual Background for Natural Attenuation

IA. Background and Purpose .........................................................................................................1

IA.1 Document Organization ..................................................................................................1

IA.2 Purpose of Document.....................................................................................................1

IA.3 Applicable Regulatory Criteria ........................................................................................2

IA.4 Policy Framework for Use of MNA..................................................................................2

IB. Relevant Distinctions in Site Characterization for MNA of Inorganic Contaminants.................4

IC. Tiered Analysis Approach to Site Characterization ..................................................................5

IC.1 Tier I................................................................................................................................6

IC.2 Tier II...............................................................................................................................7

IC.3 Tier III..............................................................................................................................7

IC.4 Tier IV .............................................................................................................................8

ID. Role of Modeling in the Tiered Analysis Approach.................................................................10

ID.1 Developing a Conceptual Model...................................................................................10

ID.2 Types of Models............................................................................................................11

ID.2.1 Simple Calculations .............................................................................................11

ID.2.2 Mass Transport Models........................................................................................12

ID.2.3 Speciation Models................................................................................................12

ID.2.4 Reaction Models ..................................................................................................12

ID.2.5 Reactive Transport Models ..................................................................................12

ID.3 Modeling and the Tiered Analysis Approach ................................................................13

ID.3.1 Tier I – Demonstration of Contaminant Removal from Ground Water.................13

ID.3.2 Tier II – Determine Mechanism and Rate of Attenuation ....................................13

ID.3.3 Tier III – Demonstrate Capacity and Stability of Removal Mechanism ...............13

ID.3.4 Tier IV – Long-Term Performance Monitoring .....................................................14

ID.4 Choosing Modeling Software........................................................................................14

ID.4.1 Public Domain vs. Commercial Software.............................................................14

ID.4.2 Sources of Software.............................................................................................14

ID.4.3 Thermodynamic Data ..........................................................................................16

ID.5 Accounting for Uncertainty............................................................................................17

ID.6 Model Calibration and Verification ................................................................................17

IE. Long-Term Performance Monitoring and Site Closure............................................................18

IE.1 Duration and Monitoring Frequency..............................................................................19

IE.2 Monitoring of Aquifer Solids..........................................................................................19

IE.3 Monitoring Types...........................................................................................................19
vi

IE.4 Monitoring Locations.....................................................................................................20

IE.5 Modification of the Performance Monitoring Plan.........................................................22

IE.6 Periodic Reassessment of Contaminant Removal Technologies..................................22

IF. References..............................................................................................................................23

Section II - Technical Basis for Natural Attenuation in Ground Water

IIA. Physical Transport Mechanisms..............................................................................................26

IIA.1 Basics of Ground-Water Flow and Solute Movement...................................................26

IIA.2 Colloidal Transport of Inorganic Contaminants.............................................................27

IIA.2.1 Implications for Monitored Natural Attenuation....................................................27

IIB. Contaminant Sorption to Aquifer Solids..................................................................................28

IIB.1 Adsorption.....................................................................................................................29

IIB.1.1 Reactive Mineral Phases Involved in Adsorption.................................................29

IIB.1.2 Surface Functional Groups on Aquifer Solids and the Impact on Surface 

Charge .................................................................................................................31

IIB.1.3 Weak and Strong Adsorption Regimes ...............................................................32

IIB.2 Precipitation ..................................................................................................................33

IIB.2.1 Precipitation from Solution...................................................................................33

IIB.2.2 Coprecipitation .....................................................................................................35

IIB.2.3 Surface Precipitation............................................................................................35

IIB.2.4 Mineral Transformation.........................................................................................36

IIB.3 Implications for Natural Attenuation Assessment .........................................................36

IIC. Microbial Impacts on Inorganic Contaminant Attenuation ......................................................36

IIC.1 Characteristics of Aquifer Microbiology ........................................................................37

IIC.2 Microbial Controls on Subsurface Redox State............................................................37

IIC.3 Impacts on Contaminant Speciation and Attenuation...................................................39

IIC.3.1 Contaminant Oxidation-reduction Reactions .......................................................39

IIC.3.2 Biosorption and Intracellular Bioaccumulation.....................................................39

IIC.3.3 Methylation and Demethylation............................................................................40

IIC.4 Implications for Natural Attenuation Assessment .........................................................40

IID. References..............................................................................................................................40

Section III - Site Characterization to Support Evaluation of MNA

IIIA.Site Hydrogeology...................................................................................................................43

IIIA.1 Characterization Objectives ..........................................................................................43

IIIA.2 Geologic Characterization.............................................................................................44

IIIA.2.1 Saturated Porous Media ......................................................................................44

IIIA.2.2 Saturated Fractured Media ..................................................................................44

IIIA.3 Hydrologic Characterization..........................................................................................45

IIIA.4 Ground-Water/Surface-Water Interactions....................................................................47

IIIA.5 Hydrogeologic Data Interpretation................................................................................47

IIIA.5.1 Attenuation Rate Estimates .................................................................................48

IIIA.5.2 Contaminant Flux ................................................................................................48

IIIA.5.3 Source Term Characteristics................................................................................50

IIIB.Contaminant Quantification, Distribution and Speciation........................................................50

IIIB.1 Aqueous Characterization Approaches ........................................................................50

IIIB.1.1 Filtration ..............................................................................................................51

IIIB.2 Solid Phase Characterization Approaches ...................................................................52

IIIB.2.1 Sampling and Fractionation .................................................................................53

IIIB.2.2 Total Amount........................................................................................................53

IIIB.2.3 Structurally Defined Form ....................................................................................54

IIIB.2.4 Operationally Defined Form.................................................................................55

IIIB.2.4.1 Sequential Extractions...............................................................................55
vii

IIIB.2.4.2 Sequential Extraction Considerations........................................................58

IIIB.2.5 Attenuation Capacity............................................................................................59

IIIB.3 Model Representations to Interpret Contaminant Sorption Observations....................61

IIIB.3.1 Distribution Coefficient/partition Coefficient, Kd ..................................................61

IIIB.3.2 The Langmuir Model............................................................................................61

IIIB.3.3 The Freundlich Isotherm......................................................................................61

IIIB.3.4 Mechanistic Models for Predicting Sorption - Surface Complexation..................62

IIIB.3.5 Mineral Solubility..................................................................................................63

IIIB.3.5.1 Coprecipitation Reactions..........................................................................64

IIIB.3.5.2 Thermodynamic Data ................................................................................65

IIIC.Characterization of System Redox and Underlying Microbial Processes...............................66

IIIC.1 Process Identification....................................................................................................66

IIIC.1.1 Redox Measurements..........................................................................................67

IIIC.2 Capacity ........................................................................................................................68

IIIC.3 Stability .........................................................................................................................69

IIIC.4 Microbial Community Characterization .........................................................................70

IIIC.4.1 Standard and Emerging Techniques....................................................................70

IIIC.4.2 Molecular Characterization ..................................................................................70

IIIC.4.3 Sampling Considerations .....................................................................................71

IIIC.5 Implications for Natural Attenuation Assessment .........................................................71

IIID.References..............................................................................................................................72
viii

Figures

Figure 1.1 Conceptual distinction between organic versus inorganic contaminant plume 

behavior where natural processes are active within the ground-water aquifer..............5

Figure 1.2 Conceptual depiction of the data collection effort to demonstrate whether sorption 

to aquifer solids attenuates contaminant transport in ground water..............................6

Figure 1.3 Example of a network design for performance monitoring, including target zones 

for monitoring effectiveness with respect to specific remedial objectives. ..................15

Figure 2.1 Conceptual view of attenuation as the interaction of the contaminant with aquifer 

constituents to form a product resulting in attenuation................................................25

Figure 2.2 Cross-sectional view of differences in solute migration due to differences in hydraulic 

conductivity with accompanying differences in ground-water velocity and the 

spreading of the solute front caused by dispersion.....................................................26

Figure 2.3. Representation of an aquifer mineral surface with (a) an outer-sphere surface 

complex; (b) an inner-sphere surface complex; (c) a multinuclear surface complex or 

a surface precipitate; and (d) absorption, or solid state diffusion and substitution of 

the sorbate in the mineral structure.............................................................................28

Figure 2.4 Examples of contaminant-specific sorption processes that may lead to attenuation 

of the ground-water plume...........................................................................................29

Figure 2.5 Diagrammatic sketch of the structure of 1:1 and 2:1 phyllosilicate minerals. ............30

Figure 2.6 Surface charge of some hydroxides from pH 2 to 10 measured in different 

electrolyte solutions shown in parentheses; positive and negative surface charge 

shown above and below the x-axis, respectively.........................................................31

Figure 3.1 Geologic block diagram and cross section depicting a stream environment. .............45

Figure 3.2 Potential effects of changes in ground-water flow direction on temporal trends in 

contaminant concentrations.........................................................................................46

Figure 3.3 Elements of a conceptual site model for monitored natural attenuation of inorganic 

contaminants. ..............................................................................................................49

Figure 3.4 Illustration of two approaches for determining attenuation rate constants within a 

contaminant plume. .....................................................................................................50

Figure 3.5 pH-dependent solubility trend of orpiment predicted using two different Gibbs free 

energy of formation values. .........................................................................................65
ix

Tables

Table 1.1 Synopsis of site characterization objective to be addressed throughout the tiered 

analysis process and potential supporting data types and/or analysis approaches 

associated with each tier...............................................................................................9

Table 1.2 Example software packages for modeling groundwater flow and mass transport .....15

Table 1.3 Example software packages for speciation in inorganic geochemical systems .........15

Table 1.4 Example software packages for modeling reactive transport in inorganic 

geochemical systems. .................................................................................................16

Table 1.5 Example internet sources of thermodynamic data useful in constructing 

geochemical models ...................................................................................................16

Table 1.6 Objectives for performance monitoring of MNA...........................................................18

Table 2.1 Important functional groups in humic substances that impact surface charging 

behavior and contaminant binding...............................................................................32

Table 2.2 Major mineral classes in aquifers and soils ...............................................................34

Table 2.3 Relationships among Q, K, and Ω .............................................................................35

Table 2.4 Range of hydrogen concentrations for a given terminal electron-accepting 

process that can be used for classification of the redox status within the 

contaminant plume .....................................................................................................38

Table 3.1 Sequential extraction procedure of Tessier et al. (1979) ............................................56

Table 3.2 Summary of reagents used to selectively dissolve iron oxides and sulfides .............57

Table 3.3 BCR extraction scheme applied to 1 gram of sample ................................................58

Table 3.4 Synopsis of the various surface complexation models (SCMs) commonly 

employed to describe solute partitioning to solid surfaces .........................................62

Table 3.5 Ground-water redox parameters and measurement approaches ..............................67

Table 3.6 Methods that may be employed for estimating the oxidation and reduction 

capacity for solid materials (from USEPA, 2002) .......................................................68

Table 3.7 Standard and emerging techniques for microbial community characterization ..........70
x

Acknowledgments

This document represents a collective work of many individuals with expertise in the policy and technical aspects of selecting and implementing cleanup remedies at sites with contaminated ground water. 

Preparation of the various components of this document was undertaken by personnel from the USEPA 

Office of Research and Development (ORD), Office of Superfund Remediation and Technology Innovation 

(OSRTI), and Office of Radiation and Indoor Air (ORIA), as well as technical experts whose participation 

was supported under USEPA Contract No. 68-C-02-092 to Dynamac Corporation, Ada, Oklahoma, through 

funds provided by ORIA and OSRTI. Contributing authors are listed below along with their affiliation:

Contributing Author Affiliation

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

74820

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

Steven Acree USEPA/ORD, National Risk Management Research Laboratory, Ada, OK 

74820

Steve Mangion USEPA/ORD/OSP, Region 1, Boston, MA 02114

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

74820

Ann Azadpour-Keeley USEPA/ORD, National Risk Management Research Laboratory, Ada, OK 

74820

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

OH 45268

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

Albuquerque, New Mexico 87185

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

Richland, WA 99352

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

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

Critical and constructive reviews were provided by Jim Weaver (USEPA/ORD National Exposure Research 

Laboratory, Athens, GA), George Redden (Idaho National Laboratory, Batelle Energy Alliance), and 

Sue Clark (Washington State University, Chemistry Department). Pat Bush (Ada, OK) is acknowledged 

for her technical editing to provide consistency in formatting and grammar. Martha Williams (Contract 

#68-W-01-032) assisted with final editing and formatting for publication. This effort is dedicated to the 

memory of Kenneth Lovelace, whose insight and patience made it a reality.
xi

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 first volume of a set of three volumes that address the technical 

basis and requirements for assessing the potential applicability of MNA as part of a ground-water remedy 

for plumes with non-radionuclide and/or radionuclide inorganic contaminants. Volume 1, titled “Technical 

Basis for Assessment,” consists of three sections that describe 1) the conceptual background for natural 

attenuation for inorganic contaminants, 2) the technical basis for attenuation of inorganic contaminants 

in ground water, and 3) approaches to site characterization to support evaluation of MNA. Emphasis is 

placed on characterization of immobilization and/or degradation processes that may control contaminant 

attenuation, as well as technical approaches to assess performance characteristics of the MNA remedy. 

A tiered analysis approach is presented to assist in organizing site characterization tasks in a manner 

designed to reduce uncertainty in remedy selection while distributing costs to address four primary issues:

1. Demonstration of active contaminant removal from ground water & dissolved plume stability; 

2. Determination of the mechanism and rate of attenuation;

3. Determination of the long-term capacity for attenuation and stability of immobilized contaminants; and 

4. Design of performance monitoring program, including defining triggers for assessing MNA 

failure, and establishing a contingency plan.

Detailed discussion is provided on the importance of acquiring site-specific data that define ground-water 

hydrogeology and chemistry, the chemical and mineralogical characteristics of aquifer solids, and the 

aqueous and solid phase chemical speciation of contaminants within the ground-water plume boundary. 

Technical distinctions are drawn between characterization efforts to evaluate the applicability of MNA 

as part of a cleanup remedy for organic versus inorganic contaminants. Emphasis is placed on the 

need to collect site-specific data supporting evaluation of the long-term stability of immobilized inorganic 

contaminants. Also included is discussion on the role of analytical models as one of the tools that may 

be employed during the site characterization process. This discussion is intended to provide context to 

contaminant-specific site characterization approaches recommended in the remaining two volumes of 

this document. 

This document is limited to evaluations performed in porous-media settings. Detailed discussion of 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 
xii

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 supersede any guidance. 

It is intended for use as a technical reference in conjunction with other documents, including OSWER 

Directive 9200.4-17P, “Use of Monitored Natural Attenuation at Superfund, RCRA Corrective Action, and 

Underground Storage Tank Sites” (http://www.epa.gov/swerust1/directiv/d9200417.pdf). 
1

Section I

Conceptual Background for Monitored Natural Attenuation

Kenneth Lovelace, Stuart Walker, Ronald Wilhelm, Robert Puls, Robert G. Ford, Richard T. 

Wilkin, Steven Acree, Steve Mangion, Patrick V. Brady, Craig Bethke

IA. Background and Purpose

IA.1 Document Organization

The purpose of this document is to provide a framework 

for assessing the potential application of monitored natural 

attenuation as part of the remedy for inorganic contaminant 

plumes in ground water. It is organized into three volumes 

that provide: Volume 1 - a general overview of the framework 

and technical requirements for application of Monitored 

Natural Attenuation (MNA); Volume 2 - contaminant-specific 

discussions addressing potential attenuation processes and 

site characterization requirements for non-radionuclides, 

and Volume 3 - contaminant-specific discussions 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 these detailed overviews 

are described below.

The non-radionuclide contaminants selected for this document include: arsenic (As), cadmium (Cd), chromium (Cr), 

copper (Cu), lead (Pb), nickel (Ni), nitrate, perchlorate, and 

selenium (Se). The selection of these contaminants by 

USEPA was based on several criteria. First, a 1994 booklet 

containing information regarding common chemicals found 

at Superfund sites throughout the nation was consulted 

(USEPA, 1994). The most commonly found inorganic contaminants were included for consideration in this document. 

Another document specific to metal-contaminated Superfund sites (USEPA, 1995) identified arsenic (As), cadmium 

(Cd), chromium (Cr), and lead (Pb) as primary contaminants 

of concern based on toxicity, industrial use, and frequency 

of occurrence at Superfund sites. Second, selection was 

based on chemical behavior considering chemical traits 

such as: toxicity, ion charge (cation vs. anion), transport 

behavior (conservative vs. non-conservative), and redox 

chemistry to cover a broad range of geochemical behavior 

(USEPA,1999a; USEPA, 1999b; USEPA, 2004). Finally, 

USEPA regional staff were asked to nominate inorganic 

contaminants that occurred frequently or that were problematic in their Regions. The above list of nine inorganic 

contaminants was selected from this process.

The radionuclide contaminants selected for this document 

include: americium (Am), cesium (Cs), iodine (I), neptunium 

(Np), plutonium (Pu), radium (Ra), radon (Rn), technetium 

Tc), thorium (Th), tritium, strontium (Sr), and uranium (U). 

The selection of these contaminants by EPA was based 

on two criteria. First, a selected element had to be one 

of high priority to the site remediation or risk assessment 

activities of the USEPA (USEPA, 1993; USEPA, 2002). 

Second, selection was based on chemical behavior considering chemical traits such as: toxicity, cations, anions, 

conservatively transported, non-conservatively transported, 

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

2004). By using these characteristics of the 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. 

IA.2 Purpose of Document

This document is intended to provide a technical resource 

for determining whether MNA is likely to be an effective 

remedial approach for inorganic contaminants1 in ground 

water. This document is intended to be used during the 

remedial investigation and feasibility study phases of a 

Superfund cleanup, or during the equivalent phases of a 

RCRA Corrective Action (facility investigation and corrective 

measures study, respectively). The decision to select MNA 

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

Superfund Record of Decision (ROD) or a RCRA Statement 

of Basis (or RCRA permit).

The USEPA expects that users of this document will include 

USEPA and State cleanup programs and their contractors, 

especially those individuals responsible for evaluating 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).



	 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

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 

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 over-

	 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.

seeing regulatory agency to identify likely characterization 

and cleanup objectives for a particular site prior to investing 

significant resources.

Use of this document is generally inappropriate in complex 

fractured bedrock or karst aquifers. In these situations the 

direction of ground water flow can not be predicted directly 

from the hydraulic gradient, and existing techniques may 

not be capable of identifying the pathway along which 

contaminated groundwater moves through the subsurface. 

Understanding the contaminant flow field in the subsurface 

is essential for a technically justified evaluation of an MNA 

remedial option. MNA will not generally be appropriate 

where site complexities preclude adequate monitoring 

(USEPA, 1999c).

Because documentation of natural attenuation requires 

detailed site characterization, the data collected can be 

used to compare the relative effectiveness of other remedial 

options and natural attenuation. The technical information 

contained in this document can be used as a point of reference to evaluate whether MNA by itself, or in conjunction 

with other remedial technologies, is sufficient to achieve 

site-specific remedial objectives.

IA.3 Applicable Regulatory Criteria 

All remedial actions at CERCLA sites must be protective 

of human health and the environment and comply with 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 §30

0.430(a)(1)(iii)(F)). In general, drinking water standards 

provide relevant and appropriate cleanup levels for ground 

waters that are a current or potential source of drinking 

water. However, drinking water standards generally are 

not relevant and appropriate for ground waters that are 

not a current or potential source of drinking water (see 

55 FR 8732, March 8, 1990). Drinking water standards 

include federal maximum contaminant levels (MCLs) and/

or non-zero maximum contaminant level goals (MCLGs) 

established under the Safe Drinking Water Act (SDWA), 

or more stringent state drinking water standards. Other 

regulations may also be ARARs as provided in CERCLA 

§121(d)(2)(B).

IA.4 Policy Framework for Use of MNA

The term “monitored natural attenuation” is used in this 

document when referring to a particular approach to 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-
3

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, the 1999 

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 protective under anticipated changes in site 

conditions. The actual policy language is given below:

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

sorption or oxidation-reduction reactions), 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 sorption

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 radionulides (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

	 When a contaminant is associated with a solid phase, it is usually not

known if the contaminant is precipitated as a three-dimensional molecular coating on the surface of the solid, adsorbed onto the surface

of the solid, absorbed into the structure of the solid, or partitioned into

organic matter. “Sorption” will be used in this Directive to describe, in

a generic sense (i.e., without regard to the precise mechanism) the

partitioning of aqueous phase constituents to a solid phase.
4

surface soil pathways be carefully evaluated and

eliminated as potential sources of external direct

radiation exposure.

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.)

The 1999 OSWER Directive provides the context for the 

Agency’s expectations for evaluating the feasibility of employing MNA as part of a cleanup remedy for contaminated 

ground water. As indicated by the sections from the Directive that are transcribed above, it also points out specific 

issues concerning what constitutes natural attenuation for 

inorganic contaminants. In practice, most of the technical experience developed to date has primarily dealt with 

evaluations of MNA as applied to remediation of organic 

contaminant plumes. While this experience provides some 

perspective for the scope of site characterization that may 

be warranted to evaluate MNA for inorganic contaminants, 

there are some important distinctions that bear on the 

types of required data and the approaches available to 

obtain these data. The following section elaborates these 

distinctions in order to provide context for the technical 

aspects relevant to MNA for inorganic contaminants and 

the steps needed to implement a technically defensible site 

characterization effort.

IB. Relevant Distinctions in Site 

Characterization for MNA of Inorganic 

Contaminants

As stated within the OSWER Directive on MNA (USEPA, 

1999c), natural attenuation processes are those that ‘reduce mass, toxicity, mobility, volume or concentration of 

contaminants’. Inorganic contaminants discussed within 

this document include both non-radioactive and radioactive constituents. For radioactive contaminants, radioactive 

decay processes result in the reduction of risk derived 

from radiation exposure. The rates of radioactive decay 

(characterized by the decay half-life) are known for the 

radioisotopes of concern, thus facilitating this aspect of site 

characterization. Guidelines for assessing the feasibility of 

MNA as a component of ground-water cleanup for radio-

	 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.

nuclides are provided in Volume 3 of this document. For 

non-radioactive inorganic contaminants and radionuclides 

possessing long decay half-lives, immobilization within the 

aquifer via sorption to aquifer solids provides the primary 

means for attenuation of the ground-water plume. In general, an inorganic contaminant can be transferred between 

solid, liquid, or gaseous phases present within the aquifer, 

but the contaminant will always be present. Contaminant 

immobilization will prevent transport to sensitive receptors 

at points of compliance. There are limited examples where 

degradation of inorganic contaminants may be a viable 

attenuation process (e.g., biological degradation of nitrate 

or perchlorate), but degradation is not a viable process for 

most of the inorganic contaminants discussed in this document. For inorganic contaminants subject to degradation 

or reductive transformation processes, the supporting site 

characterization will likely be consistent with the approach 

employed to assess MNA for organic contaminant plumes 

(e.g., USEPA, 1998; USEPA, 2001; see also specific discussions for nitrate and perchlorate in Volume 2). The following 

discussion provides context for the potential significance 

of immobilization as a means for natural attenuation of 

inorganic contaminants in ground water.

There is an important distinction between site characterization as applied to assessment of MNA for organic and 

inorganic contaminants. For organic contaminants, site 

characterization typically is focused towards determining the 

mechanism of contaminant degradation and the capacity 

of site conditions to sustain degradation for treatment of 

the mass of contaminant within the plume. This analysis 

may include identification of ground-water characteristics 

and degradation byproducts that are characteristic for contaminant degradation. Thus, much of the emphasis on site 

characterization for MNA of organic contaminants has been 

directed towards the collection and analysis of ground-water 

samples. In some cases, this characterization effort may 

have been supplemented with the analysis of contaminant 

degradation behavior through the use of microcosm experiments employing aquifer solids collected within the plume 

boundary. For inorganic contaminants in which immobilization onto aquifer solids provides the primary means for 

attenuation of the ground-water plume, characterization of 

the solid substrate within the aquifer plays a more significant 

role during site assessment. In this case risk reduction in 

ground water is realized through the sorption of the inorganic contaminant onto aquifer solids in combination with the 

long-term stability of the immobilized contaminant to resist 

remobilization due to changes in ground-water chemistry. 

The importance of this distinction between natural attenuation for organic and inorganic contaminants is emphasized 

in Figure 1.1. In essence, for inorganic contaminants one 

can consider the existence of two distinct ‘plumes’ within 

the boundary of the ground-water plume: 1) the dissolved 

or “mobile” plume (including dissolved contaminant and 

contaminant associated with mobile colloids), and 2) the 

solid phase or "immobile" plume resulting from sorption of 

the contaminant to aquifer solids (Figure 1.1). Thus, for 

inorganic contaminants there are two overriding objectives 

to address through site characterization:
5

1) Demonstration of removal of the inorganic contaminant from the dissolved phase leading to a stable 

or shrinking ground-water plume and, 

2) Demonstration of stabilization of the inorganic 

contaminant immobilized onto aquifer solids such 

that future re-mobilization will not occur to a level 

that threatens health of environmental receptors.

Evaluating the overall success of natural attenuation for 

inorganic contaminant remediation will require demonstrating that the rate and capacity for inorganic contaminant 

attenuation meets regulatory objectives and, in addition, 

that inorganic contaminant immobilization is sustainable to 

the extent that future health risks are eliminated. The latter 

requirement necessitates identifying the chemical speciation of the inorganic contaminant partitioned to the solid 

phase. This information is critical towards identifying the 

process controlling attenuation and evaluating the long-term 

stability of the immobilized contaminant relative to observed 

or anticipated changes in ground-water chemistry. 

Site characterization to support evaluation of MNA as a 

remedial alternative will involve assessment of contaminant 

transport in the aquifer. In general terms, this process will 

include assessment of ground-water hydrology and the 

biogeochemical processes that control contaminant migration within the plume. Defining the processes that control 

contaminant immobilization (or degradation) along the paths 

of ground-water flow will necessitate collection of a range 

of data that define the dynamics of system hydrology, the 

chemical characteristics of ground water, and the properties of the aquifer solids. In order to screen out sites that 

are inappropriate for selection of MNA, it is recommended 

that collection of site-specific data be conducted in stages 

that serve to minimize expenditures while providing insight 

into the potential existence of natural processes that may 

attenuate contaminant migration. Description of a tiered 

analysis approach for organizing site characterization tasks 

is provided in the following section.

IC. 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 

site-specific data are collected. The decision-making approach presented in this document includes three decision 

tiers that require progressively greater information on which 

to assess the likely effectiveness of MNA as a remedy for 

Figure 1.1 Conceptual distinction between organic versus inorganic contaminant plume behavior where natural

processes are active within the ground-water aquifer. Natural attenuation of inorganic contaminants is

viable only if the immobilized contaminant remains stable and resistant to remobilization during changes

in ground-water chemistry.
6

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.

IC.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 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. Determination of contaminant sorption onto 

aquifer solids could be supported through the collection of 

aquifer cores coincident with the locations of ground-water 

data collection and analysis of contaminant concentrations 

on the retrieved aquifer solids. Illustration of the type of 

data trend anticipated for a site where sorption actively 

attenuates contaminant transport is provided in Figure 1.2. 

The spatial distribution in aqueous and solid contaminant 

Figure 1.2 Conceptual depiction of the data collection effort to demonstrate whether sorption to aquifer solids attenuates contaminant transport in ground water. The left side of the diagram provides a cross-sectional view

of the spatial distribution of the contaminant concentration in ground water and co-located aquifer solids

for a site where sorption attenuates contaminant transport. The trend in aqueous and solid contaminant

concentrations for this scenario is depicted in Panel (A) to the right. Panel (B) depicts the relationship

between aqueous and solid contaminant concentrations for a site where sorption does not attenuate

contaminant transport.
7

concentrations for a site where sorption attenuates 

contaminant migration is depicted on the left side of the 

illustration. Anticipated relationships between aqueous 

and solid contaminant concentrations for sites with and 

without active contaminant attenuation via sorption are 

depicted on the right side of the illustration in Panels (A) 

and (B), respectively. Specifically, where sorption onto 

aquifer solids is occurring, there should be an increasing 

trend in solid phase contaminant concentrations as a 

function of increasing aqueous concentration. In contrast, 

no change in solid phase contaminant concentrations as a 

function of increasing aqueous concentration is indicative 

that attenuation is not occurring. Ultimately, sites that 

demonstrate ground-water plume expansion and a lack of 

contaminant sorption (for contaminants subject to sorption) 

would be eliminated from further consideration of MNA as 

part of the cleanup remedy.

IC.2 Tier II

The objective under Tier II analysis would be to eliminate 

sites where further analysis shows that attenuation rates 

are insufficient for attaining cleanup objectives established 

for the site within a 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 inorganic contaminants 

will typically involve calculation of the apparent transfer 

of mass from the aqueous to the solid phase, based 

on sampling of ground water and/or aquifer solids. It is 

recommended that these estimates be based as much 

as possible on field measurements rather than 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 is provided in 

Section I.D.) These seepage velocities can be combined 

with measured contaminant concentrations to estimate 

mass flux (mass per time per area) for each contaminant, 

in each hydrostratigraphic unit. The necessary data might 

include physical parameters such as hydraulic 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 

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 degradation or immobilization), and 

the chemical speciation of the contaminant in ground-water 

and/or the aquifer solids. Contaminant speciation refers to 

both oxidation state characterizations [e.g., As(III) vs. As(V)] 

as well as specific associations with chemical constituents 

in aquifer solids (e.g., precipitation of Pb carbonate vs. 

adsorption of Pb to iron oxides). Evaluations of the 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.

IC.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 either degradation or 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 
8

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.

IC.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 

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 
9

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 1.1. The types of data collected 

early in the site characterization process would typically be 

required for selection of appropriate engineered remedies, 

including characterization of the system hydrology, groundwater 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 

Table 1.1 Synopsis of site characterization objective to be addressed throughout the tiered analysis process and 

potential supporting data types and/or analysis approaches associated with each tier.

Tier Objective Potential Data Types and Analysis

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

• General ground-water chemistry data for preliminary evaluation of contaminant degradation

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
10

attenuation processes (Volume 1, Section II) and the types 

of site characterization data needed for their identification 

(Volume 1, Section III). The following discussion provides 

perspective on the role of model applications in the site 

characterization process, the types of models that might 

be employed to help meet the objectives set forth under 

each tier, and potential limitations in the availability and 

adequacy of available model codes.

ID. Role of Modeling in the Tiered Analysis 

Approach

Design of the site characterization effort and analysis of 

site-specific data in support of assessing the suitability of 

MNA as a component of the ground-water remedy is dependent on development of a Conceptual Site Model (CSM) 

that identifies site conditions and processes that influence 

contaminant transport. The CSM also provides the underpinning for selecting and developing model applications that 

provide a set of tools for evaluating transport processes, 

reaction mechanisms, attenuation capacity within the aquifer, 

and the sensitivity of the attenuation process to changes in 

site conditions. The types of models that may be employed 

as part of the site characterization process include simple 

calculations, speciation models, reaction models, transport 

models, and reactive transport models. Most modeling 

undertaken in support of an application will be quantitative, 

involving computer programs that require special skills to run 

correctly. The contaminated natural system being modeled 

is physically-, chemically-, and biologically-complex, and the 

modeler must have a thorough knowledge of the processes 

that affect the specific contaminants of concern. Site-specific 

data collected to define the physical, chemical, and biological 

characteristics of the aquifer are required to calibrate components of the analytical models and test the validity of model 

predictions. Deriving meaningful modeling results is likely to 

require expenditure of significant amounts of time, and entail 

considerable expense. This planning should occur early in 

the site assessment process, so that the modeling can be 

integrated with the evaluation of the site and the appropriate 

data can be collected.

To obtain the best results at the least expense, it is important to 

develop a valid modeling plan before beginning the modeling 

itself. Developing such a plan will likely require the combined 

talents of a group of specialists, including those familiar with 

the site and those with expertise in applying quantitative 

modeling of physical, chemical, and biological systems to 

real-world problems. This section is devoted to giving general 

perspective to the design and implementation of the modeling 

strategy. In addition to the following discussion, the reader is 

also referred to the document entitled “Documenting GroundWater Modeling at Sites Contaminated with Radioactive 

Substances” (USEPA, 1996).

ID.1 Developing a Conceptual Model

Initially, the CSM is developed based on a general knowledge 

of ground-water hydrogeology, ground-water geochemistry, 

and known properties of the specific contaminant. With 

acquisition of data that maps out the spatial and temporal 

heterogeneity of the subsurface system, the CSM can be 

updated. In general, there are more physical, chemical, and 

biological processes operating in the subsurface of any given 

site than can reasonably be accounted for in a modeling study. 

The modeling effort begins with the careful identification of the 

processes that play significant roles in contaminant migration 

and attenuation at the site. In this way a conceptual model 

emerges that will eventually be coded into the input streams 

of the software packages that will produce the modeling results. If a correct and robust conceptual model is not derived, 

the modeling results, no matter how detailed or expensive, 

will contribute little to understanding the site, and will not be 

supportive of the MNA application.

While it is important to begin modeling with a well-planned conceptual model, the conceptual model may evolve as modeling 

and collection of site-specific data proceeds. The processes of 

observation and measurement and of modeling are, in practice, 

closely interconnected. Initial observation and measurement 

suggests a conceptual model, which supports development 

of quantitative models. The results from application of these 

quantitative models, in turn suggest additional important observations and measurements, which better constrain model 

design and implementation. In this way, the conceptual model 

is updated in an iterative fashion, as progressively more is 

learned about the site. The most significant step in developing a 

conceptual model of natural attenuation at the site is to identify 

the transport and reaction mechanisms that significantly affect 

the mobility of contaminants there. Once these mechanisms 

have been identified, the logical components that will comprise 

the conceptual model can be selected.

The evaluation of transport refers to analysis of the flow of 

ground-water through the aquifer. The rate and direction of 

ground-water flow will be governed by the physical characteristics of the aquifer solids as well as the factors controlling inputs 

of water into the aquifer. Spatial and temporal heterogeneity or 

variability in these factors determines details of the mathematical construction of analytical models used to evaluate fluid and 

contaminant migration through the aquifer. In characterizing 

transport, it is important to ask questions such as:

•	 Does groundwater migrate through the bulk aquifer matrix, through fractures or heterogeneities in the matrix, or 

both?

•	 Does solute diffusion from areas of rapid flow to those 

with stagnant conditions affect contaminant transport on 

a scale finer than the envisioned numerical gridding, so 

that a dual porosity model is required?

•	 Should the medium be considered homogeneous or heterogeneous on the scale envisioned for the nodal blocks 

in the numerical gridding?

•	 Are medium properties best assigned deterministically, or 

according to a stochastic algorithm?

•	 Is hydrodynamic dispersion described well in a Fickian 

sense (i.e., in terms of dispersivity, according to Fick’s 

law), by differential advection through a numerical gridding, or in both ways?
11

•	 How can the model be calibrated to reflect as accurately 

as possible transport rates through the subsurface?

•	 What additional data need to be collected to characterize 

flow and calibrate the transport model? Such data might 

include the distribution of hydraulic head, the evolution of 

the contaminant plume through time, and the results of 

tracer tests.

Evaluation of contaminant migration in ground water relies on 

determination of the types of chemical reactions that control 

contaminant degradation or immobilization. Thus, determination of specific reaction mechanisms that may be active within 

a ground-water plume provides the basis for constructing 

analytical models employed to evaluate performance of the 

attenuation process and project contaminant transport into 

the future. To characterize the reaction mechanisms driving 

attenuation, it is necessary to ask questions such as:

•	 Does the contaminant adsorb to solid surfaces? If so, onto 

what surfaces, and as what type of surface complex? 

Does it desorb readily?

•	 Is the contaminant chemically oxidized or reduced? Is the 

reaction catalyzed by mineral surfaces, or promoted by 

microbial activity? If so, what is the catalyst or microbial 

species?

•	 Does the contaminant precipitate as a solid phase? If it 

does, what is the phase, and what is its solubility?

•	 Might complexation of the contaminant with chemical 

constituents in solution affect its mobility?

A conceptual model can be thought of as a combination of 

the logical components describing the various aspects of 

transport and reaction at a site. For example, choice of how to 

represent hydrodynamic dispersion, the equations to account 

for sorption of contaminant species onto solid surfaces, rate 

laws describing the kinetics of redox reactions, and equations 

defining rates of microbial metabolism all contribute to the 

conceptual model. Since a conceptual model is no more 

than the sum of its components, and an analytical model is 

simply the realization of a conceptual model, the final modeling results are no better than the components selected. 

ID.2 Types of Models

There are several types of models that may prove useful for 

characterizing attenuation processes at a site. In general, 

in approaching a specific question, it is most expedient to 

begin working with the simplest applicable model, adding 

complexity to the study as necessary. It is wise to avoid the 

temptation to begin by constructing the “ultimate” model, one 

that accounts for all aspects of transport and reaction at a site. 

Highly complex models are difficult to work with, expensive 

to produce, and difficult to interpret. A more efficient strategy 

is to begin with simple models of various aspects of the system, combining these as necessary into progressively more 

complex models, until reaching a satisfactory final result, one 

that reproduces the salient aspects of the system’s behavior 

without introducing unnecessary complexity.

ID.2.1 Simple Calculations

Simple calculations performed by hand or via computer applications may provide an important component to the overall 

modeling strategy. For purposes of this document, two 

modeling approaches that fall under this category include 

simplified calculation approaches to evaluate a range of 

process outcomes and specific mathematical formulas used 

to calculate input parameters needed for implementation of 

more complex transport or reaction models. An example of 

a simplified calculation approach would be the calculation of 

the mass of contaminant and the mass of reactant within a 

predefined volume of the aquifer for the purpose of assessing if sufficient reactant mass is available for an identified 

attenuation process. This type of calculation is simplified in 

the sense that one may assume that the rate of the reaction 

is unimportant. Thus, while this type of calculation provides 

a general sense of the relative degree to which the aquifer 

could support attenuation, it does not likely provide a sufficiently accurate representation of the actual efficiency of the 

attenuation process. However, the utility of this calculation 

approach is to provide some perspective as to the relative importance of investing resources to fully characterize reactant 

mass or flux. Several examples of the second category of 

this model type, specific mathematical formulas, are provided 

at the following USEPA website - http://www.epa.gov/athens/learn2model/part-two/onsite/index.html. This website 

provides on-line access to a suite of prepackaged tools (or 

“calculators”) for performing site assessment calculations. 

Several examples relevant to site characterization advocated 

within this document include:

• “Hydraulic Gradient Calculation” for assessing the 

direction(s) of ground-water flow employing head 

measurements in wells spaced horizontally across the 

site;

• “Vertical Gradients” for assessing the potential for vertical water transport within the aquifer based on head 

measurements in closely-spaced, vertically nested 

wells with identical screen lengths;

• “Vertical Gradients with Well Screen Effects” for assessing the influence of variable screen lengths in vertically 

nested wells on the calculated vertical gradient; and

• “Average Borehole Concentrations” to illustrate the potential impact on contaminant concentrations measured 

for samples collected from a single long-screened well 

in an aquifer with a depth-varying concentration and a 

depth-varying hydraulic conductivity field.

These simplified models support analysis of the adequacy 

of the location and construction of ground-water wells, 

which underpins the adequacy of the monitoring design 

to provide samples and data reflective of the site-specific 

conditions. They may also be used to provide reasonable 

estimates for parameters needed as input to more complex 

mass transport or reactive transport models. Since both 

modeling approaches provide a means for preliminary assessment of site data and potentially improving design of 

the monitoring network, they play an important role in the 

site characterization effort.
12

ID.2.2 Mass Transport Models

Mass transport models seek to describe the flow of ground 

water at a site, and the transport of chemical species within 

the flow. Mass transport models are generally implemented 

as transient simulations in one, two, or three dimensions. 

Strictly speaking, a mass transport model considers the 

migration of non-reacting species. In reality, many mass 

transport codes can consider simple reaction scenarios, such 

as partitioning of a species onto the solid surface according to a constant partitioning factor. Mass transport models 

can seldom be relied upon for describing natural attenuation, 

because they lack sophisticated knowledge of chemical and 

biological processes, but are nonetheless valuable in evaluating 

a site’s potential for MNA. The models are well developed 

and straightforward to run; they are useful tools for simulating 

the rate and pattern of groundwater flow at a site.

Mass transport modeling might be applied to figure the transit time of contaminants within the site, absent attenuating 

processes. The models find use in applying the results of 

tracer tests to calibrate the flow field. Some reactive transport 

models (described below) accept externally determined flow 

fields as input, so running a mass transport model may be a 

required preliminary to a full reactive transport model.

ID.2.3 Speciation Models

Speciation models seek to describe the distribution of chemical 

mass between solution, minerals, mineral surfaces, gases, 

and biomass. Models of this class are useful because they 

can predict the conditions under which contaminants might 

be attenuated by sequestration, and those in which they are 

likely to be mobile in the ground-water flow. For example, a 

speciation model might demonstrate that a contaminant is 

likely to adsorb to the surface of a component of the aquifer 

solids over the pH range of interest. Or, the model might 

show that the contaminant will tend to complex strongly with 

dissolved chemical species, leaving it mobile and resistant 

to attenuation.

Speciation models are implemented via the assumption that 

the modeled system is in chemical equilibrium or, more commonly, partial chemical equilibrium. A model can be configured 

to account for:

• Reactions among species in solution, including protonation-deprotonation, redox, and complexation reactions.

• Adsorption reactions onto solid surfaces, possibly 

including minerals and organic matter.

• Precipitation and dissolution reactions, to predict whether a mineral is saturated in solution, or undersaturated 

or supersaturated.

• Gas solubility reactions, to account for the dissolution 

of coexisting gases into solution, or the loss of gas 

species from solution.

Where redox reactions play a critical role in the attenuation 

reaction, it may be important to use a speciation model that 

can account for redox disequilibrium. Microbial respiration, for 

example, is driven by the transfer of electrons from donating 

to accepting chemical constituents, including the inorganic contaminant. It may be critical, therefore, to characterize the redox 

state of ground water at a site in an accurate and meaningful 

manner to fully evaluate redox-driven reactions that influence 

contaminant attenuation. Redox reactions in shallow ground 

water rarely attain a state of equilibrium (e.g., Lindberg and 

Runnells, 1984), which limits the utility of analytical models that 

describe the distribution of chemical species in ground water 

based on a single parameter such as dissolved oxygen (DO) 

concentration or Eh (e.g., as measured using a DO or platinum 

electrode, respectively). Geochemical models that describe 

redox in terms of a single parameter may be limited in their 

accuracy and/or flexibility in describing the redox characteristics 

of the ground-water system. An alternative approach to the 

model design would be to employ a flexible description of 

redox in a state of chemical disequilibrium (e.g., as discussed 

in Bethke, 1996, Chapter 6.). This type of modeling approach 

allows the user to specify for each element the mass found 

in the various possible redox states and reports the energy 

(i.e., the Nernst Eh) associated with the half reaction for each 

pairing of the element’s oxidized and reduced states.

ID.2.4 Reaction Models

Reaction models are similar to speciation models in that they 

consider the distribution of chemical mass, but have the additional ability of modeling the chemical evolution of the system. 

Like speciation models, it is commonly necessary to use a 

reaction model with a flexible description of redox disequilibrium, as well as suitable models to describe adsorption and 

precipitation reactions. Where appropriate, the model should 

be able to account for the kinetics of species sorption, redox 

reactions, mineral precipitation and dissolution, or microbial 

metabolism. Examples of the application of reaction models 

in an MNA application include:

•	 Sequestration of contaminants onto a mineral surface as 

the mineral forms, such as the complexation of heavy 

metals in mine drainage onto ferrihydrite.

•	 Precipitation of contaminant-bearing minerals, according 

to a kinetic rate law appropriate for the chemical conditions at the site.

• Immobilization of a contaminant by oxidation or reduction, according to a kinetic rate law.

•	 Biotransformation of a contaminant by microbial life, using a rate equation for fermentation or cellular respiration 

appropriate for conditions at the site.

ID.2.5 Reactive Transport Models

Reactive transport models, as the name suggests, are the 

coupling of reaction models to transport models. Unlike a 

reaction model, a reactive transport model predicts not only 

the reactions that occur in the ground-water flow, but the distribution of those reactions across the site through time. A 

reactive transport model of a site may have several advantages 

over a simple reaction model, including:

•	 The ability to account for heterogeneity at the site, such as 

an uneven distribution of a sorbing mineral, variation in 
13

pH conditions, or the differential development of microbial 

populations.

•	 The ability to describe evolution of a contaminant plume 

through space and time.

Reactive transport modeling is a relatively complex and timeconsuming undertaking, since it combines the data needs 

and uncertainties inherent in modeling reaction as well as 

transport, and because the calculation procedure may require 

a significant amount of computing time. It may be the capstone of the modeling effort, but is seldom the best tool for 

initial scoping of the attenuation capabilities at a site. Such 

modeling, on the other hand, may play an important role in the 

site characterization effort, because it represents the integration 

of all of the components of the conceptual model.

ID.3 Modeling and the Tiered Analysis 

Approach

As described in Section IC of this document, a tiered analysis approach is recommended for organizing the collection 

of site-specific data and providing a means for screening 

out sites inappropriate for selection of MNA as part of the 

ground-water remedy. Previously, possible applications of 

models of varying complexity throughout the tiered analysis 

process were provided in Table 1.1. The following discussion provides additional context for evaluating the potential 

role of model applications during the site characterization 

process. 

ID.3.1 Tier I – Demonstration of Contaminant 

Removal from Ground Water

The application of models under Tier I pertains primarily to initial characterization of hydrology and evaluating 

whether measured ground-water characteristics may support immobilization processes. Assessment of hydrology 

may include calculation of horizontal or vertical gradients 

to assess the predominant direction(s) of ground-water 

flow. This information could be used to guide installation 

of monitoring points within the aquifer for collection of 

ground-water and aquifer solids samples. Evaluation of 

contaminant immobilization potential may involve use of 

chemical data collected from ground-water and/or aquifer 

solids samples as input into a speciation model to assess 

the potential for contaminant precipitation or adsorption 

onto aquifer solids. For example, speciation calculations 

based on measurements of alkalinity and dissolved lead 

within the ground-water plume may indicate saturation or 

oversaturation with respect to precipitation of lead carbonate. Conversely, measurements of ground-water chemistry 

and extractable iron concentrations in aquifer solids could 

serve as parameter inputs into a speciation model with the 

capability of describing contaminant adsorption onto iron 

oxides. It is recommended that these latter calculations 

be used as secondary lines of evidence in support of sitespecific measurements that demonstrate active sorption of 

the contaminant onto aquifer solids within the plume.

ID.3.2 Tier II – Determine Mechanism and Rate of 

Attenuation

Modeling at this stage in the evaluation process should be 

closely integrated with observational study. In studying the 

mechanism of contaminant removal from ground water, careful attention should be paid to assuring collection of sufficient 

data to fully define the components of the conceptual model. 

For example:

•	 If a precipitating phase is identified by x-ray diffraction, 

spectroscopy, or electron microscopy, it will be necessary 

to characterize the phase’s solubility.

•	 If reaction with solid surfaces is identified as an important 

attenuation process, it will be necessary to collect sufficient data to properly parameterize an adsorption model 

that describes the specific mechanism of adsorption, as 

described in Section IIIB.

•	 It may be necessary to establish a kinetic rate law describing precipitation of the contaminant into solid phases, or 

its adsorption onto solid surfaces, where these reactions 

may occur at different rates throughout the plume due to 

the concentrations of aqueous or solid reactants. 

In determining the rate of the attenuation process, modeling 

may be used to describe chemical fluxes in the system and 

rate of species uptake or production during chemical reaction. 

Modeling might be specifically employed to estimate the time 

frame required to sequester the contamination sufficiently to 

meet cleanup objectives, where the attenuation reactions are 

kinetically controlled.

ID.3.3 Tier III – Demonstrate Capacity and 

Stability of Removal Mechanism

Model applications under Tier III would be directed toward 

assessment of the capacity of the aquifer to attenuate the 

mass of contaminant within the ground-water plume and the 

long-term stability of an immobilized contaminant. Reaction 

models and/or reactive transport models might be employed 

to evaluate the extent of contaminant removal throughout 

the plume. Use of these model types allows assessment of 

rate-dependent reactions and/or the influence of decreases 

in the flux of reactants due to changes in concentration or 

ground-water flow that might occur over time. These same 

models may be employed to evaluate ground-water conditions 

that may remobilize contaminants sorbed to aquifer solids. 

These evaluations may prove most useful for situations in 

which laboratory testing may be less practical. For example, 

model simulations may be employed to examine the stability 

of the attenuated contaminant for hypothetical situations not 

reflected in existing ambient ground water. For example, modeling might be applied for a number of specific purposes:

•	 To test the chemical feasibility of specific remobilization 

scenarios, such as infiltration of pristine groundwater, a 

shift in oxidation state (perhaps due to waterlogging), or 

a change in pH (due to soil acidification, for example).
14

•	 To figure reactant fluxes required to remobilize the contaminant.

•	 To evaluate the possible effects of chelating agents, such 

as organic acids, in the groundwater.

These model applications provide a means to project 

system behavior under conditions that do not currently 

exist, but could feasibly develop. They provide a source of 

information that further reduces the uncertainty of reliance 

on MNA as a permanent remedy.

ID.3.4 Tier IV – Long-Term Performance 

Monitoring 

Under Tier IV of the analysis process, modeling provides a 

tool for designing a long-term monitoring plan, as well as a 

contingency remedy for cases where unanticipated changes 

in site conditions leads to failure of the MNA remedy. Modeling 

tasks that might be performed at this stage include:

•	 Optimizing the location of monitoring wells for long-term 

observation.

•	 Optimizing the frequency of sample collection events 

based on knowledge of ground-water flow dynamics at 

the site.

•	 Identifying critical chemical parameters to monitor based 

on model simulations to examine the sensitivity of attenuation process rate or capacity to changes in ground-water 

composition.

•	 Identifying critical parameters to monitor based on model 

simulations to evaluate conditions leading to contaminant 

re-mobilization.

These model applications provide a means for designing the 

monitoring program to best evaluate remedy performance 

and provide site managers with a context for evaluating 

possible decreases in the efficiency of the attenuation 

process.

ID.4 Choosing Modeling Software

Once a modeling strategy has been developed and a conceptual model defined, a computer software package (or 

packages) will be needed to compute the modeling results. 

A number of software packages exist for modeling physical, 

chemical, and biological processes in natural systems. No 

single package is best for all problems; one seeks the package or packages that best satisfies the objectives of the site 

characterization process. Significantly, software packages 

designed for analyzing problems of the MNA of organic contaminants (e.g., Bioplume III; USEPA, 1997) are generally not 

suitable for studying the fate of inorganic contaminants. The 

first step in selecting software involves identification of packages incorporating features needed to evaluate the conceptual 

model. The selection process should amount to more than 

compiling a checklist of features. It is important to determine 

if the features work well for the situation in question.

It is critical to consider the efficiency of the software, not only 

in computing time, but the time required to configure each 

run and render the modeling results in a suitable graphical 

form. One should, therefore, inspect carefully the documentation from potentially suitable packages, and run test cases. 

In evaluating a commercial package, insist on inspecting 

the documentation before buying. Avoid licensing software 

without being allowed a trial period, or a period during which 

the software may be returned for a full refund.

ID.4.1 Public Domain vs. Commercial Software

Modeling software falls into two categories, public domain 

and commercial. Public domain codes can generally be downloaded over the internet or purchased for a minimal charge; 

some codes are obtained by personal request addressed to the 

developer. A public domain code has a number of potential 

advantages: there is little or no up-front cost; the source code 

is in many cases available, allowing the modeler to correct 

bugs and add features; and there may be a body of experienced users available for consultation or troubleshooting at 

minimal or no charge. A commercial code also has potential 

advantages: it may be written by a group of professional programmers; there may be people assigned to support users, 

offer training, and fix bugs; documentation may be superior; 

there is more likely to be an intuitive user interface; the code 

may be easier to use than public domain alternatives; and it 

may offer superior graphics for rendering results. In general, 

distributors of commercial codes hope they can convince 

customers that the up-front costs of their product will be offset 

in the long run by quality and savings, principally by improving the productivity of the people involved in the modeling 

process, and by speeding project completion.

ID.4.2 Sources of Software

A considerable number of software packages that can be applied to the analysis of inorganic contaminant attenuation in 

ground water are available in the public domain and from commercial sources. Tables 1.2–1.4 list examples of various types 

of commonly applied packages and their sources. Additional 

packages may be found by searching the internet, and from 

software retailers such as Rockware, Inc. (www.rockware.

com) and Scientific Software Group (www.scisoftware.com). 

New software packages appear frequently, others fall into 

disuse or are no longer supported and updated, and new 

releases of the various packages add features and fix bugs. 

As such, no attempt is made in this document to provide 

exhaustive listings of software packages applicable to 

MNA assessments, nor to judge the suitability or compile 

the features of various packages. In evaluating software, the 

reader will be well served by considering in light of his or her 

own needs only the most recent available information. The 

following discussion provides some issues to consider during selection of a software package.

Issues to consider during selection of a mass transport 

model and a representative list of commonly applied models 

(Table 1.2):

• Whether the model operates in two or three dimensions, 

or both.

• Whether the model can account for dispersion in the 

manner chosen.
15

• If the model accounts for saturated flow (flow below the 

water table), unsaturated flow (above the water table), 

or both.

• The deterministic or stochastic method or methods the 

model can use to represent heterogeneity in the properties of the medium (hydraulic conductivity, dispersivity, 

and so on) across the modeling domain.

Issues to consider during selection of a geochemical speciation 

model and a representative list of commonly applied models 

(Table 1.3):

• A flexible description of redox state. A disequilibrium 

scheme in which each redox couple can be set to its 

own redox potential is commonly required. 

• The ability to account for sorption or surface complexation in a manner appropriate for the site.

Issues to consider during selection of a reaction model, in 

addition to those relevant for a speciation model, and a representative list of commonly applied models (Table 1.3):

• An accounting for the kinetics of redox reactions, 

whether occurring in the fluid phase, catalytically on 

mineral surfaces, or promoted by enzymes.

• The ability to account for the kinetics of mineral precipitation and dissolution reactions invoked as an attenuation mechanism, using appropriate rate laws.

• A model of microbial metabolism based on valid chemical principles. The metabolic model should treat the 

Table 1.2 Example software packages for modeling groundwater flow and mass transport.

Software Source

FEFLOW Groundwater Modeling, Inc. www.ssg-int.com/

GMS Environmental Modeling Systems, Inc. www.ems-i.com/GMS/gms.html

Modflow-2000 U. S. Geological Survey water.usgs.gov/nrp/gwsoftware/modflow.html

Visual Modflow Waterloo Hydrogeology www.visual-modflow.com

GroundWater Vistas www.groundwater-vistas.com

Table 1.3 Example software packages for speciation in inorganic geochemical systems. Each of these packages 

except Wateq4F also has at least some capability for modeling reaction processes.

Software Source

Chess Ecole des Mines de Paris chess.ensmp.fr/

Eq3/6 Lawrence Livermore National Laboratory 

www.llnl.gov/IPandC/technology/software/softwaretitles/eq36.php

Mineql+ Environmental Research Software http://www.mineql.com/

MinteqA2 U.S. EPA http://www.epa.gov/ceampubl/mmedia/minteq/

Phreeq-C U.S. Geological Survey wwwbrr.cr.usgs.gov/projects/GWC coupled/phreeqc/index.html

The Geochemist’s 

Workbench® University of Illinois www.geology.uiuc.edu/Hydrogeology

Visual Minteq KTH (Sweden) www.lwr.kth.se/english/OurSoftware/Vminteq/index.htm

Wateq4F U.S. Geological Survey water.usgs.gov/software/wateq4f.html
16

metabolism as a balanced chemical reaction, accounting for not only consumption of substrate species, but 

generation of product species. The software should 

also account for how the amount of energy available 

in the environment affects metabolic rate, and for the 

growth and decay of biomass.

Issues to consider during selection of a reactive transport 

model, in addition to the points raised above about mass 

transport, speciation, and reaction models, and a representative list of commonly applied models (Table 1.4):

• Whether the model can work in one, two, or three 

dimensions.

• Compatibility of the model with the mass transport and 

reaction models chosen. For example, can the reactive 

transport model import a flow field predicted by the 

mass transport model?

• Time to solution, since reactive transport modeling can 

require considerable amounts of computing time.

ID.4.3 Thermodynamic Data

Most software packages are confıgured to accept any external database, provided that it is presented in the proper 

format. A number of databases have been compiled for 

various purposes, and many of these are available already 

formatted to be read directly into one or more of the widely 

distributed geochemical models. A list of various internet 

sites from which thermodynamic data can be downloaded 

in various formats is provided in Table 1.5. Additional databases might be located by consulting web pages and the 

latest documentation for the various geochemical modeling 

packages, and by searching the internet. Since updates to 

posted databases may be conducted infrequently, it may 

be worthwhile to verify the database incorporates currently 

accepted thermodynamic data based on a review of the 

technical literature. 

Table 1.4 Example software packages for modeling reactive transport in inorganic geochemical systems.

Software Source

Crunch Lawrence Livermore Laboratory www.csteefel.com/

HYTEC Ecole des Mines de Paris www.cig.ensmp.fr/~vanderlee/hytec/index.html

PHAST U. S. Geological Survey 

wwwbrr.cr.usgs.gov/projects/GWC coupled/phast/index.html

Phreeq-C U.S. Geological Survey 

wwwbrr.cr.usgs.gov/projects/GWC coupled/phreeqc/index.html

The Geochemist’s 

Workbench® 

Professional1

University of Illinois www.geology.uiuc.edu/Hydrogeology

1 The “Xt” package in previous releases.

Table 1.5 Example internet sources of thermodynamic data useful in constructing geochemical models.

Source URL

Ecole des Mines de Paris ctdp.ensmp.fr/

Japan Nuclear Cycle Development Institute migrationdb.inc.go.ip/

Murdoch University (Australia) Íess.murdoch.edu.au/iess/iess_home.htm

National Institute of Standards and Technology webbook.nist.gov/

Nuclear Energy Agency (France) www.nea.fr/html/dbtdb/

University of Illinois www.geology.uiuc.edu/Hydrogeology/hydro thermo.htm

University of Illinois at Chicago tigger.uic.edu/~mansoori/TRL html
17

ID.5 Accounting for Uncertainty

For a model constructed in support of an MNA application, 

there are a number of sources of uncertainty, including:

• Error in chemical analyses. The accuracy and completeness of chemical analyses vary widely. Routine 

chemical analyses performed by commercial laboratories are in many cases of insuffıcient quality to support 

geochemical and reactive transport modeling. Several 

useful checks for internal consistency are available 

in the American Water Works Association “Standard 

Methods” volume (Clesceri et al., 1998), and computer 

programs (e.g., Aq•QA, www.aqqa.com) are available 

for performing these tests automatically. Geochemical modeling applications require complete chemical 

analyses, including not only the contaminants of interest, but the major ion chemistry, pH, and distribution 

of metals among their mobile redox states.

•	 Error in determining hydrologic parameters. Measuring 

representative values of hydrologic parameters such as 

hydraulic conductivity and dispersivity can be difficult, because these values may change with the scale on which 

they are observed. Laboratory measurements, therefore, 

may give different results than well tests (e.g., slug and 

bail tests, pumping tests), which may in turn differ from 

values representative of the site as a whole. Measured 

hydrologic parameters are important, but may need to 

be calibrated to observations from the site, including 

perhaps the rate of plume advance or the migration of a 

tracer injected into the subsurface.

• Sample choice and dataset size. Significant error can 

be introduced by sampling bias, although this bias 

is not always obvious or even avoidable. Laboratory 

measurements of hydrologic properties, for example, 

are commonly made on samples that can be recovered 

intact, even though the fractured or poorly consolidated 

portions of the medium, left unsampled, control flow. 

Fluid samples may be taken from monitoring wells 

completed in highly conductive layers, where they can 

be extracted rapidly, leaving unaccounted significant 

quantities of residual contamination in slightly less conductive layers. Finally, the number of samples available 

or monitoring wells constructed is in some cases too 

small to comprise a statistically significant dataset.

•	 Incompleteness and inaccuracy of the thermodynamic 

database. To provide meaningful results, a geochemical 

or reactive transport model has to include each of the 

aqueous species, minerals, gases, and adsorbed species 

important at the site, and the data for these species need 

to be accurate. The thermodynamic databases available 

for geochemical study vary widely in breadth and accuracy.

•	 Error in model components. Each of the components 

of which the model is constructed is a potential source 

of error. Components likely to contribute to error include 

kinetic rate laws, surface complexation (sorption) models, 

and descriptions of the effects of microbial metabolism.

•	 Conceptual errors. Perhaps most significantly, model 

results can be affected by failure to conceptualize the 

problem completely and accurately. If an important 

process is not accounted for, or accounted for in an inaccurate fashion, the modeling results will likely be rendered 

useless.

The modeler accounts for uncertainty by experimenting with 

the model to discover which sources of uncertainty affect 

the results significantly. This uncertainty can subsequently 

be reduced, for example, by making new measurements or 

refining critical observations. Another source of uncertainty 

is the limited possibility to obtain measured site-specific values 

for some of the model parameters due to the complexity of 

the geochemical model. It is recommended that the results 

of uncertainty analysis be provided for the purpose of site 

decisions. This information would include the sources and 

potential ranges of all input data along with the origin of input 

data (i.e., review of technical literature, model calibration, field 

testing, or estimation).

ID.6 Model Calibration and Verification

Developing a quantitative model of contaminant attenuation in the subsurface may entail considerable uncertainty. 

Parameters needed to constrain the model are seldom 

known precisely, parameter inputs may not be available and 

require estimation, and the conceptual model itself may need 

refinement. Due to these uncertainties, it is necessary to 

calibrate the model to observations, and to verify that the 

model behaves in a manner that adequately describes the 

natural system. The processes of calibration and verification 

are closely related, since calibration brings the model into 

alignment with observed data. A model that (1) utilizes to the 

greatest extent possible parameter values specific to the site, 

and (2) is calibrated to the observed evolution and distribution 

of the contaminant plume, therefore, is most likely to be readily 

verified. It is recommended that steps taken to calibrate the 

model application be documented and provided for review in 

order to build confidence in the use of this assessment tool.

Model verification requires that the model predict an independent set of observations, i.e., a set separate from those 

used for calibration. For example, a model that predicts the 

attenuation of chromate by chemical reduction might be “fit” 

on the basis of a plume or section thereof, and subsequently 

used to predict the behavior of another plume at the same 

site. The initial fitting would presumably involve arriving at 

reasonably precise estimates of the most uncertain inputs – in 

this case reduction rates, electron donor loads, and so on. If 

the subsequent independent prediction accurately reflects field 

observations, this result would lend credence to the model. 

Here, “accurate reflection” of field predictions probably means 

predicting correctly the speed at which the plume is retreating 

and estimating the rate of overall contaminant mass reduction 

to within a factor no greater than five. Predictions that do not 

achieve this level of accuracy should prompt further refinement of the model.

This discussion has been intended to point out that models 

may serve as a useful tool that can be employed as part of 

the evaluation process for selection of MNA as a remedy. 
18

However, the complexity of the modeling effort and the 

potential level of uncertainty associated with model predictions indicate that pursuit of more direct lines of evidence 

is critical to the tiered analysis process. The acquisition of 

these data will depend on establishing a network of monitoring locations throughout the aquifer. The site-specific 

data collected from these monitoring locations provide 

the means to identify the attenuation process and assess 

the performance characteristics of the MNA remedy. As 

with any technology used as part of a cleanup remedy, 

continued assessment of remedy performance is critical 

for ensuring attainment of cleanup goals. The following 

discussion provides context for the eventual design of the 

performance monitoring program leading to site closure for 

situations in which MNA provides a viable component to 

the ground-water remedy.

IE. Long-Term Performance Monitoring and 

Site Closure

The performance of the MNA remedy must be monitored to 

determine compliance with site-specific remedial objectives 

identified in remedy decisions. This long-term monitoring 

is often the largest expenditure incurred in the course of 

cleanup and, for this reason alone, should be considered 

at the earliest stages of remedial investigation. Because 

the time horizons for successful implementation of an 

MNA remedy are often expected to be long, it is critical 

that particular attention is paid to long-term monitoring 

plans. Detailed discussions of the performance monitoring 

framework and monitoring plan development have recently 

been published (USEPA, 2003). Although that discussion 

focuses on attenuation of common organic contaminants, 

the framework and many of the principles governing plan 

development are also applicable to inorganic constituents. 

However, there are conceptual differences with respect to 

the outcome of the MNA remedy for inorganic contaminants. With the exception of situations where degradation 

reactions transform harmful contaminants (e.g., nitrate 

or perchlorate) into innocuous constituents, contaminant 

mass is not reduced during MNA for inorganic contaminants. The MNA process results in relocation, dispersion, 

and ultimately chemical conversion of the original source 

zone. Therefore, the purposes of performance monitoring 

are to demonstrate degradation to innocuous materials 

and immobilization of contaminants. It is recommended 

that site closure be considered only after degradation and

immobilization within the risk level specified in the remedy 

decision are demonstrated and shown to have long term 

stability.

Development of a performance monitoring plan is site 

specific in nature. Monitoring objectives and quantifiable 

performance criteria are developed to evaluate temporal 

and spatial remedy performance with respect to the sitespecific remedial action objectives. Much of the monitoring 

to demonstrate performance of the MNA remedy will fall 

into three basic categories: 1) ambient monitoring to assess 

background contaminant levels and the status of relevant 

ambient geochemical indicators (e.g., EH, pH); 2) process 

monitoring to assure the progress of chemical attenuation; 

and 3) monitoring to detect plume expansion. 

Within this framework, the OSWER Directive 9200.4-17P 

(USEPA, 1999c) provides eight specific objectives to be met 

by the performance monitoring program of an MNA remedy 

(Table 1.6). The objectives usually will be met by implementing a performance monitoring program that measures 

contaminant concentrations, geochemical parameters, and 

hydrologic parameters (e.g., hydraulic gradients). Much of 

the monitoring will be focussed on ground water. However, 

periodic monitoring of aquifer solids, through soil coring, will 

be warranted in most situations. These data will be used 

to evaluate the chemical behaviour of the contaminant in 

the subsurface over time, including:

•	 Changes in three-dimensional plume boundaries,

•	 Changes in the redox state that may indicate changes 

in the rate and extent of natural attenuation, 

•	 Reduction in the capacity of aquifer materials for contaminant immobilization, and

•	 Mobile contaminant mass and concentration reductions 

indicative of progress toward contaminant removal 

objectives. 

Contaminant behavior can then be evaluated to judge the 

effectiveness of the MNA remedy and the adequacy of the 

monitoring program. 

Table 1.6 Objectives for performance monitoring of MNA (USEPA, 1999c).

1) Demonstrate that natural attenuation is occurring according to expectations,

2) Detect changes in environmental conditions (e.g., hydrogeologic, geochemical, microbiological, or other 

changes) that may reduce the efficacy of any of the natural attenuation processes,

3) Identify any potentially toxic and/or mobile transformation products,

4) Verify that the plume(s) is not expanding down gradient, laterally or vertically,

5) Verify no unacceptable impact to down gradient receptors,

6) Detect new releases of contaminants to the environment that could impact the effectiveness of the natural 

attenuation remedy,

7) Demonstrate the efficacy of institutional controls that were put in place to protect potential receptors, and 

8) Verify attainment of remediation objectives.
19

IE.1 Duration and Monitoring Frequency

As stated in the OSWER Directive (USEPA, 1999c), performance monitoring should continue until remediation 

objectives have been achieved, and longer if necessary to 

verify that the site no longer poses a threat to human health 

or the environment. Typically, monitoring is continued for 

a specified period after remediation objectives have been 

achieved to ensure that concentration levels are stable and 

remain below target levels. In order to demonstrate stability, 

verification of the achievement of target levels under conditions where the aquifer geochemistry has reestablished 

a chemical steady state with respect to ambient groundwater geochemistry will be needed. The magnitude of the 

chemical gradient between the impacted and non-impacted 

portions of the aquifer provides a reference point for evaluating establishment of steady-state conditions. A monitoring 

strategy to verify the attainment of remedial objectives and 

provide for termination of monitoring and site closure generally should be formulated during the development of the 

performance monitoring plan and updated, as necessary, 

prior to implementation. 

Monitoring frequency should be specified in the performance monitoring plan. In addition, the plan may specify 

an approach and technical criteria that could be used to 

increase or reduce the frequency as conditions change. 

Such criteria would scale monitoring frequency to match 

MNA performance and the level of understanding and 

confidence in the conditions that control attenuation at a 

given site. The most appropriate frequency for groundwater sampling is site specific and depends on several 

factors including:

•	 The rate at which contaminant concentrations may 

change due to ground-water flow and natural attenuation processes,

•	 The degree to which the causes of this variability are 

known, 

•	 The types of evaluations to be performed and the importance of the type of data in question, and 

•	 The location(s) of possible receptors relative to the 

plume.

In addition, the most appropriate frequency may vary in 

different areas of the site based on site-specific conditions 

and the intended use of the data. Similar principles are 

applied in determining the most appropriate frequency for 

sampling of aquifer solids. 

With respect to the initial frequency of ground-water sampling under the performance monitoring program, quarterly 

monitoring may often be an appropriate frequency to establish baseline conditions over a period of time sufficient 

to observe seasonal trends, responses to recharge, and 

to confirm attenuation rates for key contaminants. Quarterly monitoring for several years provides baseline data 

to determine trends at new monitoring points and test key 

hypotheses of the conceptual site model. 

More frequent monitoring of ground-water elevations may 

be warranted, particularly during the establishment of baseline conditions, to improve the characterization of groundwater flow patterns. In addition, more frequent monitoring 

may be needed to observe changes in ground-water flow 

patterns in response to other site activities, such as the 

start or cessation of ground-water extraction in off-site water 

supply wells, source control activities, and other significant 

changes in the hydrologic system.

IE.2 Monitoring of Aquifer Solids

The aquifer material may serve as the reactive media to 

which many inorganic contaminants become partitioned 

and immobilized. Therefore, periodic re-assessment of the 

capacity of aquifer materials for contaminant immobilization, 

including immobilization of radioactive contaminants and 

any harmful products of radioactive decay, often is a critical 

step in performance monitoring. There are three aspects to 

this solid-phase characterization to be addressed through 

collection of field data and laboratory testing: 

• Determination of the chemical process(es) resulting in 

contaminant immobilization, 

• Determination of the capacity of the un-reacted aquifer 

material for contaminant immobilization, and 

• Determination of the stability of the reacted aquifer 

material with respect to contaminant release. 

Characterization of aquifer material requires collection of 

core material within the existing contaminant plume and 

down gradient and side gradient to the plume. Characterization within the existing plume is used to identify the 

immobilization process(es) and capacity, while down gradient and side gradient characterization is used to re-assess 

the potential and capacity for immobilization in the event 

of plume expansion. In general, this characterization involves identification of the aquifer mineralogy to determine 

the abundance and spatial distribution of reactive solid 

component(s) and the distribution of the contaminant among 

the identified components. 

The spatial extent and density of sampling points will be 

dictated by the degree of heterogeneity of the aquifer 

material both within and outside of the existing plume 

boundary. The frequency of sampling will be dictated by 

the rate of the immobilization process with respect to fluid 

transport and the dynamics of fluid flow and chemistry. In 

general, sampling frequency will be greater within the plume 

boundary where immobilization is active. The frequency 

of sampling outside of the plume boundary will be dictated 

by the proximity of receptors and the time frame for reaching remedial objectives relative to the rate of weathering 

processes that may change the composition or mineralogy 

of the aquifer material.

IE.3 Monitoring Types

The majority of the monitoring performed to determine the 

effectiveness of the MNA remedy may be classified under 

three general headings: 
20

• Monitoring of conditions outside of the plume boundaries (ambient monitoring),

• Monitoring of natural attenuation processes (process 

monitoring), and 

• Monitoring to detect plume expansion and verify the 

lack of impact to receptors (migration monitoring).

Other types of monitoring include periodic evaluations of 

the effectiveness of any institutional controls specified in 

the remedy decision documents and, ultimately, verification 

of the attainment of all remedial objectives.

Ambient monitoring should be performed outside of the 

boundaries (e.g., hydraulically up gradient, side gradient, 

and down gradient) of the contaminant plume. The purpose 

of this monitoring is to establish background conditions and 

to provide an indication of the potential for additional plume 

migration in situations where redox state and the capacity of aquifer materials for contaminant immobilization are 

dominant controls on migration. The extent and duration of 

ambient monitoring will be influenced by the sensitivity of 

aquifer chemistry to changes in recharge water quality and 

processes that may change its composition.

Process monitoring is used to verify that attenuation is 

occurring according to prediction. If process monitoring 

indicates that attenuation is not occurring as expected, a 

change in cleanup approach may be warranted. Process 

monitoring is contaminant-specific and might include, 

for example, measurement of ground-water redox state 

or pH to assure the existence of conditions favorable for 

natural attenuation via reduction-oxidation processes or 

pH-dependent sorption as well as the monitoring of contaminants. Process monitoring parameters are discussed 

in the contaminant-specific sections in Volumes 2 and 3 of 

this document. Process monitoring should also take into 

account any impacts of ongoing or prior active treatment on 

subsequent ambient attenuation processes. For example, 

such impacts may include gradual shifts in system redox as 

water levels and/or electron donor/acceptor levels change 

after, respectively, pump and treat or in situ bioremediation 

have been halted. 

Monitoring to detect plume expansion (migration monitoring) 

and any impacts to receptors is another important aspect 

of the performance monitoring program. This monitoring objective may be met through multi-level monitoring 

performed at or near the side gradient and down gradient 

plume boundaries, beneath the plume, and near any other 

compliance boundaries specified in remedy decision documents in conjunction with monitoring of possible receptor 

locations (e.g., potable water wells or locations of ecological 

receptors) to directly verify the lack of impacts. Monitoring 

locations between the plume and compliance boundaries 

or possible receptors should be close enough to the plume 

that a contingency plan can be implemented before the contaminant can move past the point of compliance or impact 

receptors. Identifying locations for monitoring wells designed to detect migration ultimately relies on a site-specific 

assessment of contaminant migration and fate. Additional 

insight may be obtained from site-specific transport model 

predictions, where model use is conducted iteratively with 

the site characterization process so that model predictions 

are both tested and influence future data collection.

IE.4 Monitoring Locations

At many sites, the performance monitoring program will 

be three-dimensional in nature due in large measure to 

the effects of site-specific hydrogeology on contaminant 

migration. Typical target zones for monitoring a contaminant 

plume (Figure 1.3) include:

•	 Original source areas - within and immediately down 

gradient of source areas (Process Monitoring)

The monitoring objectives include the detection of any 

further contaminant releases to ground water that may 

occur and demonstration of reductions in contaminant concentrations in ground water over time. In situations where 

the original source is contained, increased contamination 

or new contaminants could be indicative of containment 

system failure.

•	 Transmissive zones with highest contaminant concentrations or hydraulic conductivity (Process Monitoring)

A change in conditions in these zones, such as an increase 

in contaminant mass, change in redox state, increased 

ground-water velocity, or exceedance of the aquifer capacity for immobilization, may lead to relatively rapid plume 

expansion.

•	 Distal or fringe portions within the plume (Process and 

Migration Monitoring)

These are areas where reduction of contaminant concentrations in ground water to levels required by remedial action 

objectives may be attained most rapidly or where plume 

expansion may be observed most readily.

•	 Outside the plume, including areas near plume boundaries and other compliance boundaries (Migration 

Monitoring)

Multi-level monitoring points, reflecting vertical differences 

in subsurface conditions, generally will be warranted at the 

side gradient, down gradient, and vertical plume boundaries; 

between these boundaries and possible receptors; and at 

any other compliance boundaries specified in remedy decision documents. Monitoring of receptor locations should 

also be included to directly verify that no impacts occur.

•	 Zones in which contaminant reductions in ground water 

appear to be less than predicted (Process Monitoring)

These are the areas where attaining cleanup standards 

within time frames specified in the remedy decision documents may be impeded due to site conditions (e.g., higher 

than anticipated concentrations of residual source materials, 

redox conditions, or exceedance of the capacity for immobilization). Such areas, if present, will be delineated through 

evaluation of data obtained throughout the performance 
21

monitoring period. These areas may require additional 

characterization to determine if additional remedial actions 

are necessary to reduce contaminant concentrations to 

desired levels.

•	 Areas representative of uncontaminated settings (Ambient Monitoring)

Sampling locations for monitoring the redox state and immobilization capacity of aquifer materials include points that 

are adjacent to but outside the plume. Data from these 

monitoring locations will often be needed to assess the 

continuation of favorable conditions for attenuation. Since 

assumptions concerning the redox state and attenuation 

capacity affect interpretation of data from the plume, such 

assumptions should be periodically evaluated like other 

aspects of the conceptual site model. Therefore, multiple 

monitoring points generally should be used to determine 

the variability of these parameters outside the plume.

Figure 1.3 Example of a network design for performance monitoring, including target zones for monitoring effectiveness with respect to specific remedial objectives.
22

•	 Areas supporting the monitoring of site hydrology

At some sites, monitoring of ground-water elevations 

at locations additional to those used for the monitoring 

of chemical parameters may be needed to determine if 

changes in ground-water flow rates and directions are occurring. Appropriate locations for placing piezometers will 

often include positions that are up gradient, side gradient, 

and down gradient of the contaminant plume, as well as 

in zones above and below the plume and near surface 

water bodies. 

IE.5 Modification of the Performance 

Monitoring Plan

The monitoring plan should be a dynamic document that is 

modified as conditions change or the conceptual site model 

is revised to reflect new information. Decisions regarding 

remedy effectiveness and the adequacy of the monitoring 

program will generally result in either:

•	 Continuation of the monitoring program without modification;

•	 Modification of the monitoring program; 

•	 Implementation of a contingency or alternative remedy; 

or

•	 Verification that remedial objectives have been met with 

subsequent termination of the monitoring program. 

Continuation of the program without modification would be 

supported by contaminant concentrations behaving according to remedial expectations while ground-water flow and 

geochemical parameters remain within ranges indicative of 

continued contaminant immobilization. Modification of the 

program, including increases or decreases in monitoring 

parameters, frequency, or locations, may be warranted to 

reflect changing conditions or improved understanding of 

natural attenuation processes at the site. In addition, modification generally would be warranted whenever remedy 

modifications are implemented, such as implementation 

of additional source removal or hydraulic control for plume 

migration. 

In situations where hydrologic and geochemical parameters 

are stable and the contaminant concentrations in ground 

water are decreasing as predicted, reductions in sampling 

frequency (e.g., semi-annual, annual, or less frequent) will 

often be warranted for process monitoring. For example, 

five years of quarterly monitoring showing predictable 

decreases in mobile contaminant concentrations might be 

the basis for decreasing the frequency to a semi-annual 

or annual basis at some sites. Ten years of semi-annual 

or annual monitoring that shows predictable decreases in 

mobile contaminant mass might likewise be the basis for 

additional decreases in frequency, depending on site conditions. Conversely, unexpected increases or lack of predicted 

decreases in contaminant concentrations may trigger additional characterization to determine the reasons for the 

behavior, increased monitoring of pertinent parameters, 

re-evaluation of the conceptual site model, and, potentially, 

the implementation of a contingent or alternative remedy. 

Changes in the frequency of monitoring to detect plume 

expansion may also be warranted as process monitoring is 

modified. However, the frequency of such monitoring should 

not be decreased to the point where insufficient time would 

be available for implementation of an effective contingency 

plan in the event of MNA remedy failure. 

Criteria for modifying the monitoring program, including the 

type and amount of data needed to support the evaluation, 

should be discussed and agreed to by stakeholders. Sitespecific criteria should be developed to define conditions 

that indicate the appropriateness of increased or decreased 

monitoring, additional characterization, re-evaluation of the 

conceptual site model, implementation of a contingency 

or alternative remedy, and termination of performance 

monitoring.

Another reason for altering the monitoring program is the 

development of more advanced monitoring technologies. 

Because long-term monitoring costs are substantial, every 

advantage of technological advances in monitoring efficiencies should be considered. This might best be done 

by assessing monitoring technology every 3 to 5 years to 

identify “off-the-shelf” monitoring approaches/equipment 

that can improve accuracy and lower costs. National 

technology verification programs are often a good source 

of such information. 

IE.6 Periodic Reassessment of Contaminant 

Removal Technologies

In addition to the routine monitoring of MNA remedy performance, it is recommended that periodic consideration 

be given to any technological advances in the efficiencies 

of source removal for inorganic contaminants. Implementation of more efficient technologies may result in reductions in the time frames for performance monitoring with 

associated reductions in cost as well as improvements 

in performance. Many sites may benefit from a Periodic 

Remedial Technology Assessment (PRTA) conducted at 

regular intervals (e.g., 5 years) throughout the performance 

monitoring program. The PRTA should consist of a rigorous 

literature search and engineering assessment of the field 

implementation of new technologies. It should involve a 

survey of cleanup efficiencies achieved by new technologies 

at sites similar to the one under consideration. The survey 

should rely on the results of national or state technology 

verification programs (e.g., USEPA Environmental Technology Verification Program, www.epa.gov/etv/; Interstate 

Technology & Regulatory Council, www.itrcweb.org). The 

PRTA should either indicate the absence of more suitable 

alternatives or suggest a faster path to site closure. The 

criteria for technology selection should be clearly stated 

during the development of the evaluation plan. The goal 

of this review should be identification of technologies that 

have a very high probability of achieving at least order-ofmagnitude reductions in contaminant mass and/or achievement of MCLs in ground water by means acceptable to 

stakeholders. A reasonable metric should be successful 

implementation of the technology as judged by impartial 

bench marking criteria at several sites where site closure 

has been achieved.
23

1F. References

Bethke, C.M. Geochemical Reaction Modeling, Concepts

and Applications, Oxford University Press, New York, 

NY (1996).

Clesceri, L.S., A.E. Greenberg, and A.D. Eaton. Standard

Methods for the Examination of Water and Wastewater, 20th Edition, American Public Health Association, 

Washington DC (1998).

USEPA. Environmental Characteristics of EPA, NRC, and

DOE Sites Contaminated with Radioactive Substances,

EPA 402-R-93-011, Office Radiation and Indoor Air, 

Washington DC (1993).

USEPA. Common Chemicals Found at Superfund Sites,

EPA/OSWER No. 9203, Office of Solid Waste and 

Emergency Response, Washington DC (1994).

USEPA. Contaminants and Remedial Options at Selected

Metal-Contaminated Sites, EPA/540/R-95/512, Washington DC (1995).

USEPA. Documenting Ground-Water Modeling at Sites

Contaminated with Radioactive Substances, EPA/540/

R-96/003, Office of Radiation and Indoor Air, Office of 

Solid Waste and Emergency Response, Washington 

DC (1996). (http:// www.epa.gov/radiation/docs/cleanup/540-r-96-003.pdf)

USEPA. Technical Protocol for Evaluating Natural Attenuation of Chlorinated Solvents in Ground Water,

EPA/600/R-98/128, Office of Research and Development, Washington DC (1998). (http://www.epa.gov/ada/

download/reports/protocol.pdf)

USEPA. UnderstandingVariation in Partition Coefficient, Kd,

Values: Volume I - Kd Model, Measurement Methods,

and Application of Chemical Reaction Codes, EPA 

402-R-99-004A, Office of Radiation and Indoor Air, 

Washington DC (1999a).

USEPA. Understanding Variation in Partition Coefficient,

Kd, Values: Volume II - Geochemisry and Available

Kd Values for Selected Inorganic Contaminants, EPA 

402-R-99-004B, Office of Radiation and Indoor Air, 

Washington DC (1999b).

USEPA. Use of Monitored Natural Attenuation at Superfund,

RCRA Corrective Action, and Underground Storage

Tank Sites, EPA/OSWER No. 9200.4-17P, Office of 

Solid Waste and Emergency Response, Washington 

DC (1999c).

USEPA. Evaluation of the Protocol for Natural Attenuation

of Chlorinated Solvents: Case Study at the Twin Cities

Army Ammunition Plant, EPA/600/R-01/025, Office of 

Research and Development, Washington DC (2001). 

(http://www.epa.gov/ada/download/reports/epa_600_

r01_025.pdf)

USEPA. Common Radionuclides Found at Superfund Sites,

EPA 540/R-00-004, Office of Radiation and Indoor Air, 

Washington DC (2002a). 

USEPA. Workshop on Monitoring Oxidation-Reduction

Processes for Ground-water Restoration, EPA/600/

R-02/002, Office of Research and Development, 

Washington DC (2002b).

USEPA. Understanding Variation in Partition Coefficient,

Kd, Values: Volume III - Review of Geochemistry and

Available Kd Values for Americium, Arsenic, Curium,

Iodine, Neptunium, Radium, and Technetium, EPA 

402-R-99-004C, Office of Radiation and Indoor Air, 

Washington DC (2004).
24
25

Section II

Technical Basis for Natural Attenuation in Ground Water

Richard T. Wilkin, Steven Acree, Steve Mangion, Patrick V. Brady, Robert G. Ford, 

Robert W. Puls, Paul M. Bertsch, Douglas B. Kent, Ann Azadpour-Keeley, 

James E. Amonette, Craig Bethke

In determining whether MNA is applicable to a site, the 

properties of the site and the properties of the contaminant 

are analyzed in order to identify the specific process (or 

processes) causing contaminant attenuation. Inorganic 

contaminant transport in the subsurface will be governed by 

the site-specific characteristics of ground-water flow and the 

chemical interactions between the contaminant and aquifer 

solids along the path (or paths) of fluid flow. The overall 

extent of contaminant attenuation will be governed by the 

velocity of ground-water flow relative to the rates of chemical 

reactions that attenuate contaminant transport. The types 

and rates of chemical reactions that result in contaminant 

attenuation will be controlled by the availability of constituents within the aquifer that interact with the contaminant 

in a manner that results in contaminant immobilization or 

transformation. In simple terms, one can view the product 

of the reaction between the contaminant and aquifer constituents (or reactants) as the specific form of attenuation. 

Examples of possible attenuation reactions are provided 

in Figure 2.1 to illustrate this conceptual viewpoint and 

the types of processes that may be active at a given site. 

The reactants in the attenuation process may be present 

in dissolved form or associated with aquifer solids (e.g., 

aquifer minerals or microbes). Thus, assessment for MNA 

will necessitate collection of site-specific data that define 

the processes controlling contaminant transport. In order to 

provide context for the types of data that may be required to 

address this assessment objective, this section will provide 

a review of the physical and biogeochemical processes that 

govern contaminant transport in ground water.

Figure 2.1 Conceptual view of attenuation as the interaction of the contaminant with aquifer constituents to form a

product resulting in attenuation. Subscript designations “(GW)” and “(S)” indicate, respectively, whether

the reactant(s) and product(s) are in ground water or associated with immobile aquifer solids. For the

degradation reaction, N2 and CO2 are likely present as dissolved gases; “goethite” is the name of a commonly occurring iron oxyhydroxide mineral in aquifers.
26

IIA. Physical Transport Mechanisms

IIA.1 Basics of Ground-Water Flow and Solute 

Movement

Ground water is rarely static and moves from areas of 

ground-water recharge (i.e., high hydraulic head) to areas 

of discharge (i.e., low hydraulic head). In a porous medium, 

ground-water flow generally obeys Darcy’s law with velocity 

proportional to the hydraulic conductivity and hydraulic gradient. In such settings, the average interstitial flow velocity, 

or seepage velocity (Vs), is the rate at which water moves 

through the pore spaces of the medium. Under natural 

conditions, ground-water movement is relatively slow with 

rates ranging from less than a foot per year to several feet 

per day (USEPA, 1991). Seepage velocities for the various 

aquifer materials at a site may often be estimated using:

Vs = Ki/ne

where,

K = the hydraulic conductivity of the medium

i = the magnitude of the hydraulic gradient

ne

 = the effective porosity or fraction of the medium 

occupied by interconnected pore space.

In general, the seepage velocity at most sites is expected 

to vary spatially due to heterogeneity in aquifer material 

properties and temporally due to fluctuations in hydraulic 

gradients.

The dominant processes that result in subsurface solute 

movement in this dynamic environment are advection and 

dispersion. Advection is the movement of a solute with the 

bulk movement of ground water (Freeze and Cherry, 1979) 

and generally is the primary solute transport mechanism 

at sites with moderate to high ground-water flow rates. 

Mechanical dispersion results in mixing of ground water 

during advection reducing dissolved solute concentrations 

within the plume and causing some solute molecules to 

travel faster and some to travel slower than the average 

ground-water velocity. A more detailed discussion of dispersion is provided by Gelhar (1993) and Gelhar et al. (1992). 

The result at the macro scale is that a solute will spread 

to occupy a larger portion of the flow field in the direction 

of ground-water flow and in transverse (perpendicular) 

directions than that due solely to advection. However, 

limitations in the ability to obtain direct measurements of 

dispersion relegate its determination primarily as a fitting 

parameter during calibration of a ground-water flow model 

for site-specific applications. At the field scale, stratification 

(with associated differences in the hydraulic conductivity of 

geologic materials) and fluctuations in hydraulic gradient 

often result in much greater differences in the movement 

of solutes relative to estimates based on the average 

ground-water flow velocity (Figure 2.2). Thus, relative to 

site-specific evaluation of contaminant transport, assessment of the degree of variability in the distribution of hydraulic conductivity within the aquifer and time-dependent 

changes in the magnitude and direction(s) of the hydraulic 

gradient most likely plays a more critical role under the site 

characterization effort. 

Figure 2.2 Cross-sectional view of differences in solute migration due to differences in hydraulic conductivity with

accompanying differences in ground-water velocity and the spreading of the solute front caused by dispersion (Keeley, 1989).
27

IIA.2 Colloidal Transport of Inorganic 

Contaminants

The association of contaminants with suspended colloidal 

material in ground water is a possible transport mechanism 

and a complicating factor for accurate estimations of the 

natural attenuation of contaminants in subsurface systems. 

The mobile colloidal phase must be highly reactive, of sufficient quantity, and stable for periods of time (i.e., remain 

in suspension due to physical or chemical perturbations to 

the aquifer) to enable the transport of a significant mass of 

contaminants. Research to date indicates colloidal facilitated transport of contaminants in ground waters, surface 

waters and in the unsaturated zone. Evidence suggests 

that colloidal transport of contaminants may be significant 

for some species under some hydrogeological conditions. 

It is important therefore for sampling methods, transport 

models, and site assessments to consider and be sensitive 

to this transport mechanism.

Colloids are generally considered to be particles with 

diameters less than 10 microns (Stumm and Morgan, 1981). 

Based on theoretical considerations, Yao et al. (1971) and 

O’Melia (1980) have estimated that the most mobile colloidal 

particles in filtration studies in porous media may range from 

0.1 to 1.0 μm. These include both organic and inorganic 

materials. Recent estimates of colloidal concentrations 

in ground water range as high as 63 mg L-1

 (Buddemeier 

and Hunt, 1988), 60 mg L-1 (Ryan and Gschwend, 1990), 

and 20 mg L-1 (Puls and Eychaner, 1990). In addition 

to a high surface area per unit mass, colloidal particles 

such as organic carbon particles, clay minerals and iron 

oxides are also extremely reactive sorbents for inorganic 

contaminants. If mobile in subsurface systems, these 

colloids can effect the migration of contaminants for much 

larger distances than many transport models would predict, 

because sorbing contaminants spend a significant fraction 

of time associated with mobile rather than immobile solids, 

and because colloid transport can be primarily along large 

diameter, fast flow paths.

Colloidal material may be released from the soil or geologic 

matrix and transported large distances given favorable hydrological and geochemical conditions. Changes in solution 

chemistry resulting from environmental pollution or changes 

in ground water recharge chemistry can bring about 

changes in the aqueous saturation state in the subsurface 

leading to precipitation of new colloid-sized inorganic solids 

that are entrained within flowing ground water, or can cause 

the dissolution of matrix cementing agents, promoting the 

release of colloid-sized particles. In addition, changes in 

the concentrations of solutes that affect colloid surface 

charge, such as pH or organic anions, can change the 

stability of colloids. An excellent review of the mechanisms 

of colloidal release, transport, and stability was published 

by McCarthy and Zachara (1989). Reference to field and 

laboratory studies provide examples of situations representative of sites with subsurface contamination where periods 

of colloid mobilization may exist. For example, Gschwend 

and Reynolds (1987) demonstrated that submicron ferrous 

phosphate colloids were suspended and presumably mobile 

in a sand and gravel aquifer. The colloids were formed from 

sewage-derived phosphate and iron released from aquifer 

solids due to reduction and dissolution of ferric iron from the 

soil. Thompson et al. (2006) have also demonstrated that 

microbially-driven redox cycling of iron bound to soils may 

also lead to release of colloidal solids via indirect impacts 

on water chemistry. In this instance, the mobilized colloidal 

fraction was dominated by organic carbon solids. 

However, field observations also point to the transience 

of colloid mobilization in the subsurface. Nightingale and 

Bianchi (1977) observed that ground-water turbidity may 

increase for periods of time due to mobilization of colloidal 

solids coincident with time-varying recharge events. These 

observed turbidity increases were abated with time as the 

aquifer returned to steady-state conditions. In addition, 

Baumann et al. (2006) observed high colloid concentrations in leachate from landfills, but colloid concentrations 

decreased rapidly in ground water down gradient from the 

landfill. Their observations suggested that the change of 

hydrochemical conditions at the interface, from a reducing, high ionic strength environment inside of the disposal 

sites to an oxidizing, low ionic strength environment in the 

ground water (together with physical filtration effects for the 

larger particles) was an effective chemical barrier for colloid 

migration. Thus, it appears that while colloid mobilization 

(along with associated contaminants) probably does occur, colloid migration is not likely to serve as a dominant 

mechanism for contaminant migration encountered at 

contaminated sites. 

IIA.2.1 Implications for Natural Attenuation 

Assessment

Many studies have documented that colloidal transport can 

occur under some hydrogeological conditions. As noted 

above, very few studies have shown that transport of contaminants via colloidal transport in the subsurface accounts 

for the predominant mass of mobile contaminant at a site. 

Where colloidal transport may be significant from a human 

health standpoint is with highly toxic elements. From a site 

characterization perspective, assessing the significance 

of colloid-facilitated contaminant transport will depend on 

the adequacy of well installation materials and construction approaches, as well as the approaches to sampling 

ground water. Since well installation results in disturbance 

of the subsurface solids, initial development to removed 

fine-grained, disturbed aquifer materials from within and 

adjacent to the well screen is a critical step prior to initiating 

retrieval of ground-water samples. In addition, stabilization 

criteria employed as indicators of the retrieval of representative ground-water samples may need to be adjusted if it 

is important to rule out colloid transport as a contaminant 

migration method. For this situation, more strict limits may 

need to be placed on turbidity stabilization observations, 

resulting in longer pumping times and/or significantly lower 

pumping rates (e.g., Jensen and Christensen, 1999). 

In addition, it is recommended that the chemical composition of the colloidal materials be identified at sites where 

field data suggests colloid-facilitated contaminant transport. 
28

This step is warranted to rule out possible sources of sample 

contamination or artifacts resulting from sample collection 

or processing. For example, the presence of iron oxide colloids may be observed for reduced ground-water samples 

containing elevated concentrations of ferrous iron [e.g., >5 

ppm Fe(II)] that are exposed to oxygen during collection 

or handling. In this case, the likely source of colloidal iron 

oxides is due to rapid oxidation and precipitation of ferrous 

iron after the ground water has been pumped from the 

subsurface. In general, the rigid and complex procedures 

needed to insure sample quality for identification of groundwater colloids may be difficult to implement within the overall 

characterization effort (e.g., Dai et al., 2002). 

Ultimately, the resources expended to this effort need to 

be balanced with the effort to characterize site hydrology 

as well as the subsurface processes that may control 

contaminant degradation or sorption to aquifer solids. In 

order to provide context to the types of data that may 

be needed to identify specific sorption processes active 

within the plume, the following sections provide detail on 

the types of mechanisms that may result in contaminant 

immobilization.

As previously outlined in Section I, the processes leading 

to contaminant attenuation may include those that cause 

reduction of contaminant mass (i.e., degradation and 

radioactive decay) or cause immobilization of the contaminant via sorption to aquifer solids. For a majority of 

the inorganic contaminants encountered at contaminated 

sites, some form of immobilization will likely dominate the 

attenuation process. The following discussion will provide 

detail on the types of sorption processes that may result in 

contaminant attenuation. This discussion will illustrate the 

factors or parameters that are a component of the sorption 

reaction in order to provide context to the types of measurements needed to support identification of the site-specific 

attenuation process and its performance characteristics 

(discussed below in Section III of this volume). In addition, 

the impact that microbial processes exert on the subsurface 

geochemistry will be discussed in order to provide context 

to factors that may dictate site-specific conditions within a 

contaminant plume. 

IIB. Contaminant Sorption to Aquifer Solids

The primary, and in most instances, the only process involved in the natural attenuation of inorganic contaminants 

is through partitioning to the solid phase. The process of 

contaminant transfer from the aqueous to the solid phase is 

generally referred to as sorption and involves three primary 

mechanisms (Sposito, 1986): adsorption, which is the accumulation of matter at the interface between the aqueous 

phase and a solid adsorbent without the development of 

a three-dimensional molecular arrangement; precipitation,

which is the growth of a solid phase exhibiting a molecular 

unit that repeats itself in three dimensions; and absorption, 

which is the diffusion of an aqueous or adsorbed chemical 

species into a solid phase (Figure 2.3). From the standpoint of monitored natural attenuation, the mechanisms 

of most relevance will be dependent on contaminant and 

site specific characteristics, such as the surface reactivity, 

solubility, and redox sensitivity of the contaminant, as well 

as the type and abundance of reactive mineral phases and 

the ground-water chemistry. 

More detailed discussion is provided below on the various 

sorption processes introduced in the preceding paragraph. 

As a point of reference, representative examples of sorption processes for specific contaminants are provided in 

Figure 2.4.

Figure 2.3 Representation of an aquifer mineral surface with (a) an outer-sphere surface complex; (b) an innersphere surface complex; (c) a multinuclear surface complex or a surface precipitate; and (d) absorption,

or solid state diffusion and substitution of the sorbate in the mineral structure (after Sposito, 1984).
29

IIB.1 Adsorption

Adsorption processes are typically categorized by the relative binding strength of interaction between the adsorbate 

(species in solution) and the adsorbent (aquifer solid). 

There is a range of binding strength for contaminant adsorption that depends on characteristics of the adsorbate, 

sorbent, and ground-water chemistry. However, discussions 

of binding strength are generally couched in terms of “weak” 

or “strong” adsorption processes, albeit a common convention in chemistry would categorize both the covalent and 

electrostatic interactions involved in adsorption as ‘strong’ 

intermolecular forces (Israelachvili ,1994). One microscopic 

distinction borrowed from the characterization of soluble 

ion pairs that is commonly used to delineate weak and 

strong adsorption involves the solvation properties of the 

adsorbate (Westall, 1986; Stumm, 1992). If solvating water 

molecules are interposed between the cation or anion and 

the surface, the adsorption complex is referred to as outer

sphere and is considered to be weak. Conversely, if upon 

adsorption the adsorbate loses waters of hydration such 

that there are no water molecules interposed between the 

cation or anion and the surface, the adsorption complex is 

referred to as inner sphere and is considered to be strong 

(Sposito, 1984). The propensity of a cation or anion to form 

either an inner-sphere or outer-sphere surface complex is 

a function of the adsorbate, the surface functional groups 

of the adsorbent, and aqueous phase chemistry (e.g., pH 

and ionic strength). 

IIB.1.1 Reactive Mineral Phases Involved in 

Adsorption

Important adsorbent phases commonly found in the 

environment include phyllosilicate minerals, metal 

oxyhydroxide phases, sulfide phases, and natural organic 

matter (Dixon and Schulze, 2002). Many phyllosilicate 

minerals possess a permanent negative charge as a result 

of the substitution of lower valence cations, i.e., Mg(II), 

Fe(II), Li(I) for Al(III) in the octahedral layer and/or Al(III) 

for Si(IV) in the tetrahedral layer (referred to as isomorphic 

substitution). There are two main classes of phyllosilicate 

minerals based on layer structure (Figure 2.5). The 1:1 

mineral layer type is comprised of one Si tetrahedral layer 

and one Al octahedral layer, which in soils and aquifers 

is commonly represented by the mineral kaolinite having 

the general formula [Si4]Al4O10(OH)8•nH2O. Kaolinite and 

related minerals generally have insignificant degrees of 

cation substitution within their octahedral and tetrahedral 

layers, and, thus generally posses a very low permanent 

negative charge. The 2:1 mineral type is comprised of one 

Al octahedral layer interposed between two Si tetrahedral 

layers comparable to the mica structures (Figure 2.5). The 

2:1 layer class is represented by a variety of minerals, 

which are classified based on the location (tetrahedral 

vs. octahedral layer) and relative amount of isomorphic 

substitution. The three major mineral classes within the 2:1 

layer type are illite (Mx

[Si6.8Al1.2](Al3Fe0.25Mg0.75)O20(OH)4), 

vermiculite (Mx[Si7Al](Al3Fe0.5Mg0.5)O20(OH)4), and smectite 

(Mx[Si8]Al3.2Fe0.2Mg0.6O20(OH)4), which display different levels 

Figure 2.4 Examples of contaminant-specific sorption processes that may lead to attenuation of the ground-water

plume. Color coding is employed to distinguish the contaminant (red), aqueous or solid phase reactants

(blue), and the product (yellow) of the reaction leading to contaminant attenuation. Absorption is illustrated as a possible sequential process that follows the adsorption of a contaminant onto a mineral component within aquifer solids.
30

of cation substitution in their tetrahedral and octahedral 

layers. The permanent negative charge imparted to 2:1 clay 

minerals by isomorphic substitution is typically balanced 

through exchange reactions involving major cations in 

ground water (e.g., Na+, K+, Ca2+, or Mg2+; represented by 

“Mx” in the formulas listed above). 

Contaminant sorption to phyllosilicates may occur via ion 

exchange or surface complexation with surface functional 

groups (see examples labeled “surface complex” and “ion 

exchange” in Figure 2.5). Due to differences in the levels 

of isomorphic substitution for the 1:1 and 2:1 clay mineral 

classes, ion exchange is usually only significant for 2:1 phyllosilicates. In addition to siloxane oxygen atoms along the 

basal plane, phyllosilicates possess two types of terminal 

ionizable OH groups, aluminol and silanol, protruding from 

the edge surface. These edge OH groups can form both 

inner- and outer-sphere complexes with metal cations and 

oxyanions depending on the pH of the bathing solution and 

on the specific characteristics of the cation or oxyanion 

(represented as “surface complex” in Figure 2.5). 

The most important surface reactive phases for both cationic and anionic contaminants in many soil and subsurface 

systems are the metal oxyhydroxide phases. These phases 

are characterized by hexagonal or cubic close-packed O or 

OH anions with Fe2+,3+,Al3+, and/or Mn3+,4+ occupying octahedral sites. These oxides are present as discrete phases 

and as complex mineral assemblages, being co-associated 

with phyllosilicates and primary minerals as coatings or 

with humic macromolecules. In soils and sediments the 

crystallinity of these phases typically varies from poorly 

ordered to well crystalline forms and grain size from the 

nanometer to micrometer scale. Among the most common 

Fe-oxyhydroxide phases found in soils and sediments are 

the poorly ordered phase ferrihydrite (Fe2

O3

.

nH2O), and the 

moderate to well crystalline phases, goethite (α−FeOOH), 

and hematite (α-Fe2O3). The most common Al oxyhydroxide 

phase found in soils and sediments is gibbsite (γ-Al(OH)3). 

Additionally, poorly ordered aluminosilicates can be important reactive phases in certain soils and these include 

the very poorly ordered allophanes (Si/Al ratios 1:2 to 1:1) 

and the paracyrstalline phase, imogolite (SiO2

.

Al2O3

.

2H2O). 

While Mn oxyhydroxides are less prevalent than Fe- and 

Al-oxyhydroxides in soils and sediments they are very important phases in terms of surface mediated redox reactions 

and because of their propensity for high metal sorption. 

The mineralogy of Mn is complicated by the range in MnO bond lengths resulting from extensive substitution of of 

Mn2+ and Mn3+ for Mn(IV). Thus, there exists a continuous 

series of stable and metastable compositions from MnO to 

MnO2 forming a large variety of minerals. Among the more 

common Mn-oxyhydroxides are pyrolusite (β-MnO2), the 

hollandite-cryptomelane family (α-MnO2), todorokite, and 

birnessite (σ-MnO2).

Figure 2.5 Diagrammatic sketch of the structure of 1:1 and 2:1 phyllosilicate minerals. Also shown are hypothetical

sorption reactions for zinc and cesium (ion exchange represented as ‘Mx’ in structural formulas for 2:1 

phyllosilicates shown in text).
31

IIB.1.2 Surface Functional Groups on Aquifer 

Solids and the Impact on Surface Charge

The reactive surface functional group for all of the metal 

oxyhydroxide phases is the inorganic OH moiety exposed 

on the outer periphery of these minerals. The reactivity of 

a specific metal oxyhydroxide is dependent on the surface 

area (SA), surface-site density (NS), the degree of coordination of the OH group to the bulk structure, and the point of 

zero charge (PZC). The charge on the mineral surface may 

impose either attractive or repulsive contributions to the 

overall adsorption reaction, depending on the type of charge 

possessed by the adsorbate. The properties of the sorbent 

that impact adsorption are controlled by both the grain size 

and specific structure of the oxyhydroxide phase. 

The surface charge of oxyhydroxide minerals and edge 

sites on phyllosilicates is derived from the protonation and 

ionization of exposed surface hydroxyl groups, represented 

by ≡ SOHn

n-1, where S represents the structural metal cation 

(e.g., Fe, Al, Mn) over a stoichiometric range from n= 0, 1, 

or 2. Thus, as a function of pH, the surface functional groups 

can be generally described with the following idealized nomenclature: ≡ SOH2

+, ≡ SOH0

, and ≡ SO-. The exact charge 

associated with the various surface functional groups is 

difficult to measure, so the main purpose of employing this 

nomenclature is to illustrate that surface charge varies as 

a function of ground-water chemistry. The gradual change 

in surface charging with pH for some common minerals is 

illustrated in Figure 2.6 and a discussion of surface site 

charging is provided below.

Natural organic matter comprised of complex polymers 

called humic substances, represents another very important reactive phase in aquifer solids. A variety of functional 

groups are present in humic substances, and, like OH 

functional groups of the inorganic metal oxyhydroxides, 

these also are characterized by pH dependent charging 

mechanisms. The primary functional groups associated with 

humic substances in terms of surface charge are carboxyl 

and phenolic groups, however the less abundant amino, 

imidazole, and sulfhydryl groups may play an important 

role in the sorption of certain contaminant metals when 

present at trace levels (Table 2.1).

Based on the previous discussion, it is apparent that the 

charge on aquifer solids can be grouped into two classes 

associated with the mechanisms that give rise to electrical 

charge associated with mineral surfaces or with organic 

functional groups. These two classes are commonly referred to as permanent (or constant) charge and variable 

charge.

• Constant charge - Constant charge is the predominant 

charge in phyllosilicate clays. Because, for the most 

part, these isomorphic substitutions occur during 

mineral formation, this charge deficit is fixed in the lattice structure and is hence unaffected by changes in 

electrolyte concentration or pH of the soil solution.

• Variable (pH dependent) charge - Variable charge is 

the predominant charge for oxyhydroxide minerals 

such as hematite, goethite, and gibbsite, as well as 

for humic substances. The metal ions in the vicinity of 

the surface of metal oxyhydroxide minerals are coordinatatively unsaturated, i.e. they are lewis acids, and 

coordinate with water molecules, which subsequently 

dissociate a proton leading to a surface layer of metal 

hydroxide functional groups. This process also occurs 

at the edges of phyllosilicate clays giving rise to SiOH 

(silanol) and AlOH (aluminol) functional groups. These 

surface hydroxyl groups can become positively or 

negatively charged by binding or dissociating a proton 

(i.e., protonation-deprotonation reactions):

S—OH + H+  S—OH2

+

S—OH  S—O- + H+

Thus, the prevalent surface charge in aquifer solids will 

be dependent on the pH of ground water and the types 

and concentrations of ions that balance the permanent 

charge association with phyllosilicates. The extent to which 

Figure 2.6 Surface charge of some hydroxides from

pH 2 to 10 measured in different electrolyte

solutions shown in parentheses; positive and

negative surface charge shown above and

below the x-axis, respectively. Ferrihydrite

[Fe(OH)3

.

nH2O] (0.001 M NaNO3) from Hsi

and Langmuir (1985); gibbsite [Al(OH)3] and

silica gel [SiO2

.

nH2O] (1.0 M CsCl) based on

Greenland and Mott (1978); goethite [αFeOOH] (0.005 M CsCl) based on Greenland and Mott (1978) (see also Hsi, 1981);

birnessite (σ-MnO2] (0.001 M NaNO3) based

on Catts and Langmuir (1986).
32

protonation or deprotonation occurs is also a function of the 

metal ion and the local binding environment of the metal 

hydroxide surface group. The highly electropositive Si4+ in 

silanols precludes the protonation of the surface hydroxyl 

and this functional group can only dissociate a proton 

under pH conditions generally encountered in ground water. 

Aluminols, on the other hand, can be either positively or 

negatively charged. Various types of hydroxyls of differing 

reactivity have been identified spectroscopically at the 

surface of metal oxides.

The charge of aquifer minerals is always electrically balanced by interactions with ions of an opposite charge 

(counter-ions). We can define two broad classes of weak 

and strong interactions (outer and inner-sphere) that act to 

neutralize the charge developed at soil mineral surfaces.

IIB.1.3 Weak and Strong Adsorption Regimes

Weak adsorption regime - Within the weak adsorption regime, simple ion exchange is the most common mechanism 

and involves the electrostatic attraction of an ionic species 

by a negative or positive charge emanating from a mineral 

surface or from functional groups associated with humic 

substances (Sposito, 1981). Long before the structures of 

reactive soil minerals were determined, it was observed 

that, under certain circumstances, there was a reversible 

and stoichiometric (based on charge) replacement of major 

cations in soils equilibrated with concentrated neutral salt 

solutions according to the reaction: 

Na2

-Xs + Ca2+

aq  Ca-Xs

 + 2Na+

aq

Soil and sediment materials are typically characterized by 

their cation exchange capacity (CEC), which is defined as 

the quantity of cations reversibly adsorbed per unit weight 

of mineral and typically expressed as cmol kg-1. The cation 

exchange capacity of 2:1 phyllosilicate clays tends to be 

constant over a wide pH range, since ionizable edge groups 

are relatively minor on a surface area basis compared to 

the permanent charge associated with planar sites. For 

2:1 phyllosilicates, cations hydrated to differing degrees 

are located in the interlayer space and can be displaced 

by other competing cations through ion exchange reactions 

(see Figure 2.5). In principle, cation exchange reactions 

involve both inner and outer sphere complexation with planar sites, although except for the special case discussed 

below for large weakly hydrated monovalent cations, such 

as K+ and Cs+, both are readily reversible. Both inner and 

outer sphere complexes can also form with O functional 

groups associated with organic macromolecules and O and 

OH atoms associated with metal oxyhydroxides, but only 

the outer sphere complexes are considered weak adsorption. The major difference between phyllosilicates having 

substantial isomorphic substitution and metal oxyhydroxides 

and humic substances, is that the CEC is highly pH dependent, increasing with increasing pH. Since reactive mineral 

phases in soils and sediments are a composed of complex 

assemblages of phyllosilicates, oxyhydroxides, and humic 

substances, CEC is always a pH dependent property. 

Strong adsorption regime - As discussed above, simple ion 

exchange is the predominant adsorption mechanism for 

phyllosilicate clays. A major exception to this is the innersphere sorption of larger unhydrated cations, such as K+

and Cs+ to oxygen atoms of two opposing siloxane ditrigonal 

cavities of collapsed layers of weathered micaceous minerals, such as illite, which can be classified as an ‘irreversible’ 

adsorption or as an absorption process. 

At this point it is important to discuss the concept of the 

reversibility of adsorption. From the perspective of chemical 

thermodynamics, the definition of a ‘reversible’ process is 

one where the initial state of the system can be restored 

with no observable effects in the system and its surroundings (Holman, 1980). The use of the term ‘irreversible’ 

from the standpoint of adsorption mechanisms is relative 

and does not strictly adhere to the thermodynamic (or 

chemical) definition in all cases. The fixation of Cs+ in illitic 

minerals is conceptually thought to proceed via an initial 

ion exchange reaction followed by an interlayer collapse 

(fast) or through the slower migration into interlayer sites 

in collapsed layers (absorption). Once Cs+ is fixed within 

the interlayer of the clay mineral, its release is not readily 

reversible via displacement with competing solutes, i.e., 

through ion exchange mechanisms. Thus, the release of 

fixed Cs+ is subsequently controlled by a process such as 

mineral dissolution. In this sense, the original fixation process is irreversible, since contaminant release would result 

from a mechanism other than the reversal of the original 

adsorption mechanism. 

The formation of a chemical bond between an adsorbate 

and a functional group on the adsorbent also falls within 

the category of a strong adsorption regime. In general, 

breaking chemical bonds requires more energy than overcoming electrostatic interactions. Metal adsorption to OH 

functional groups of oxyhydroxide phases through surface 

complexation can be illustrated by the following surface 

reaction (Stumm, 1992; McBride, 2000):

=Fe-OH-1/2 + M(H2

O)6

n+ ⇒ =Fe-O-M(H2

O)5

(n-3/2)+ + H3

O+

Table 2.1 Important functional groups in humic substances that impact surface charging behavior and contaminant binding.

Functional Group Structural Formula

Amino -NH3

Carboxyl

Carbonyl

Alcoholic hydroxyl -OH

Phenolic hydroxyl

Imidazole

Sulfhydryl -SH
33

Thus, specific adsorption of cations increases the positive 

charge at the mineral surface when n > 1, which is generally the case for transition elements, and results in the net 

release of H+ ions to the soil solution. Adsorption of anions 

from solution occurs by ligand exchange of a OH or H2O at 

the surface functional group according the following reaction 

(solution anion represented by An-): 

=Fe-OH + An- ⇒ S-A(n-1)- + OHAnion adsorption is favored by low pH, which leads to 

protonation of surface functional groups and makes them 

better leaving groups in the ligand exchange reaction.

IIB.2 Precipitation

Mineral-water reactions occur as ground water moves 

through porous media. These reactions may result in the 

removal of aquifer components due to mineral dissolution 

or result in the buildup to oversaturation and consequent 

precipitation of secondary minerals. As an outcome of 

mineral-water reactions along a flow path, fluid compositions and the mineralogical makeup of the solid phase will 

continuously evolve towards a stable or equilibrium state. 

Mineral precipitation processes in aquifer systems are an 

important group of immobilization mechanisms for inorganic 

contaminants in ground water. 

Full treatment of precipitation processes, including coverage of relevant thermodynamic and kinetic concepts, is 

outside the scope of this document. The reader is referred 

to numerous standard textbooks in geochemistry, soil science, and aquatic chemistry (e.g., Lindsay, 1979; Stumm, 

and Morgan, 1981; Drever, 1982; Sposito, 1989; Stumm 

1992; Morel and Hering, 1993; McBride, 1994; Sparks, 

1995; Langmuir, 1997; Lasaga, 1999). The purpose of this 

section is to introduce key concepts and issues regarding 

the potential impact precipitation reactions may exert on 

contaminate attenuation. In general, mineral precipitation 

in relation to contaminant immobilization can be discussed 

in the context of four widely studied processes:

• Precipitation from solution: Nucleation and growth of 

a solid phase exhibiting a molecular unit that repeats 

itself in three dimensions. Homogeneous nucleation 

occurs from bulk solution and heterogeneous nucleation 

occurs on the surfaces of organic or mineral particles. 

Heterogeneous nucleation is thought to be more important in natural systems that are rich in reactive inorganic 

and biological surfaces. Precipitation may result in the 

formation of sparingly soluble hydroxides, carbonates, 

and, in anoxic systems, sulfides. Many precipitation 

reactions have a strong dependence on pH.

• Coprecipitation: Incorporation of an element as a trace 

or minor constituent within a precipitating phase. In this 

case, the contaminant substitutes for a more concentrated component in the crystal lattice (isomorphous 

substitution). This process is distinct from adsorption 

due to incorporation of the contaminant within the bulk 

structure of the major mineral phase. Examples of 

coprecipitation include Cr(III) in hydrous ferric oxide, 

Cd(II) in calcium carbonate, and As(III) in iron sulfide.

• Surface precipitation: A precipitation process intermediate between surface complexation and precipitation 

from bulk solution. Surface precipitation represents the 

continuous growth of particles formed via heterogeneous nucleation. Macroscopic studies of adsorption 

of some solutes, particularly di-valent and tri-valent 

cations, suggest that precipitation occurs at surfaces 

under conditions where the solid is apparently undersaturated based on solution concentrations (Dzombak 

and Morel, 1990).

• Mineral transformation: Adsorbed contaminants can 

become incorporated into minerals that form as a result 

of recrystallization or mineral transformation processes 

in soils and sediments. Transformation reactions may 

be accelerated or retarded by the contaminant, and in 

some cases mineral transformation may result in the 

exclusion of the impurity contaminant from the solid 

phase. Examples include incorporation of anions, such 

as As(V), into hydrous ferric oxide and transformation 

to Fe oxyhydroxides (e.g., Ford, 2002), coprecipitation 

of metals with iron monosulfide and transformation to 

iron disulfide (e.g., Lowers et al., 2007), and layered 

double hydroxides (typically with Al) as intermediates 

between adsorbed/surface precipitated metal ions like 

Ni and Zn, and metal-ion-containing aluminosilicates 

(e.g., Ford, 2007).

The relative importance of these processes will be determined by contaminant characteristics as well as site-specific 

characteristics of the plume ground-water chemistry and 

aquifer solids. These individual processes are discussed 

in more detail below.

IIB.2.1 Precipitation from Solution

Solution precipitation or crystallization can be divided 

into two main processes: nucleation and crystal growth. 

Nucleation occurs prior to growth of a mineral crystal. Both 

nucleation and growth processes require a system to be 

oversaturated in the new phase. The probability of nucleation occurring increases exponentially with the degree of 

oversaturation. Nucleation of a new phase is often facilitated in the presence of a surface (heterogeneous nucleation) compared to bulk solution (homogeneous nucleation). 

Because nucleation and growth are processes that compete 

for dissolved solutes, at high degrees of oversaturation the 

rate of nucleation may be so fast that all excess solute is 

partitioned into crystal nuclei. In contrast, lower levels of 

oversaturation can result in the growth of existing crystals 

without nucleation. Well-formed or euhedral crystals typically develop slowly via growth from solution at low degrees 

of oversaturation. During crystal growth various chemical 

reactions can occur at the surface of the growing mineral, 

such as adsorption, ion exchange, diffusion, and formation of surface precipitates. In general, the rate of crystal 

growth is controlled either by transport of solutes to the 

growing surface (i.e., transport controlled), by reactions at 

the surface (i.e., surface controlled), or a combination of 

these factors.
34

For the most abundant cations present in aquifers and 

soils, such as Al, Si, Fe, Mn, Ca, and Mg, precipitation of 

mineral forms is common and will in many cases control 

concentrations observed in solution. At contaminated sites, 

concentrations of ground-water contaminants are typically 

several orders of magnitude below the concentrations of the 

dominant solutes in water. At low concentrations, sorption, 

surface precipitation, or formation of a dilute solid solution 

(coprecipitation) may be the more probable removal processes for contaminant species (McBride, 1994). However, 

precipitation of iron-bearing or aluminum-bearing minerals, 

for example, can have an important affect on the transport 

and fate of metal and metalloid contaminants. Major mineral 

precipitate classes encountered in aquifers and soils are 

listed in Table 2.2.

The tendency for a system to support a specific precipitation 

or dissolution reaction can be evaluated through comparison of the equilibrium solubility constant for a given solid 

phase to the ion activity product calculated using groundwater chemical data. The relative magnitude of the values 

of the equilibrium solubility product and the calculated ion 

activity product provides a measure of the saturation state of 

ground water relative to mineral precipitation or dissolution. 

A conventional method for expressing the ground-water 

saturation state is by calculation of the saturation index, 

SI, which is given by (Stumm and Morgan, 1981):

SI = ∆Gr

0

/RT + ln Q = ln Q/Kr

where ∆Gr

0

 is the standard state free energy change of the 

reaction, R is the gas constant, T is temperature in degrees 

Kelvin, Q is the reaction quotient (or ion activity product), 

and Kr is the temperature- and pressure-dependent equilibrium constant of a reaction. Another term used frequently 

in place of the saturation index is the relative saturation, 

Ω = Q/Kr. At chemical equilibrium, ∆Gr

0 = 0, Q = Kr

, and 

Ω = 1. In this special case, the solution of interest is in 

equilibrium with the mineral and no dissolution or precipitation should take place. Where ∆Gr

0< 0, the mineral cannot 

precipitate from solution and the thermodynamic driving 

force is such that mineral dissolution should occur. Where 

∆Gr

0 > 0, the mineral will likely precipitate if there are no 

limiting kinetic factors (Table 2.3).

Table 2.2 Major mineral classes in aquifers and soils.

Mineral Class Primary Mineral Contaminant Precipitate

Hydroxides

Al(OH)3

, gibbsite

Fe(OH)3, hydrous ferric oxide

FeO(OH), goethite

FeO(OH), lepidocrocite

Cu(OH)2

Cr(OH)3

Zn(OH)2

Oxides

Fe3

O4, magnetite

Fe2O3, hematite

MnO2, pyrolusite

SiO2, quartz

UO2

, uraninite

Carbonates

CaCO3

, calcite/aragonite

FeCO3, siderite

MnCO3, rhodochrosite

CdCO3, otavite

ZnCO3, smithsonite

PbCO3, cerussite

Sulfates

BaSO4

, barite

CaSO4

.

2H2O, gypsum

PbSO4, anglesite

RaSO4

Sulfides FeS, mackinawite

FeS2, pyrite/marcasite

PbS, galena

NiS, millerite

HgS, cinnabar

ZnS, sphalerite

Phyllosilicates

Al4

(OH)8Si4O10, kaolinite

K1.5Al2(OH)2Si2.5Al1.5O10, illite

Ni3Si2O5(OH)4, nepouite

Na0.3Zn3(Si,Al)4O10(OH)2 

. 4H2

O, sauconite
35

The ion activity product is a useful probe to evaluate the 

potential for contaminant precipitation. Some caution, 

however, is recommended in interpreting solution indicators 

as evidence for the presence of a particular precipitated 

solid within the plume. The observation that an ion activity 

product is equal to a corresponding solubility product is not 

unequivocal evidence that a given phase is at equilibrium 

or even present in the system (Sposito, 1984). Similarly, 

an ion activity product that is greater than a corresponding 

solubility product cannot be taken as confirmation that precipitation is occurring. To understand the state of a system 

with respect to precipitation and dissolution, it is recommended that the presence of the relevant solid phases be 

evaluated in addition to measuring the concentrations of 

solutes that participate in a precipitation reaction suspected 

to occur within the ground-water plume.

Table 2.3 Relationships among Q, K, and Ω.

Process

Saturation 

Index, 

log (Q/Kr)

Relative 

Saturation, 

Ω

Q, Kr

Mineral 

dissolution Negative < 1 Q < Kr

Mineral 

precipitation Positive > 1 Q > Kr

Equilibrium 0 1 Q = Kr

IIB.2.2 Coprecipitation

Contaminant plumes are often characterized by concentrations of dissolved solids in excess of that found in ambient 

ground water. These elevated dissolved solids may be 

derived as a component of the contaminant source or due 

to the dissolution of soil or aquifer solids during plume 

transport. Examples of these processes include interactions of acid wastes with aquifer solids leading to dissolution 

of aquifer minerals (e.g., carbonates or oxyhydroxides) or 

the development of reducing conditions driven by microbial 

degradation of organic contaminants that result in reductive 

dissolution of iron-bearing minerals. With downgradient 

transport, changes in ground-water chemistry or interaction 

with unimpacted aquifer solids may lead to precipitation 

of these major ground-water constituents out of solution. 

Contaminants may be removed from ground water at the 

location where precipitation of these major ground-water 

constituents occurs. This process is called coprecipitation, 

since the contaminant is sequestered within a newly formed 

precipitate, but only as a trace structural component within 

the precipitate. Examples of major precipitate classes with 

a coprecipitated contaminant include oxyhydroxides (e.g. 

Fe1-xCrx

(OH)3(s)), carbonates (e.g., Ca1-xCdxCO3(s)), sulfides 

(e.g., Fe1-xNixS2(s)), and phyllosilicates. The contaminant 

may be coprecipitated in a cationic or anionic form depending on ground-water chemistry and the nature of the 

precipitating phase.

For coprecipitates (or solid solutions) the concentration of 

the contaminant in ground water in contact with a precipitate 

may be reduced significantly below that observed for ground 

water in which the concentration of the contact is governed 

by the solubility of a precipitate in which the contaminant is 

a major structural component (e.g., Ca1-xCux

CO3(s), coprecipitate vs. CuCO3(s), pure precipitate). For example, the 

partial molal Gibbs free energy of a binary mixture can be 

expressed as the sum of two components: a mechanical 

mixing term and a free energy of mixing term (∆Gmixture):

G X G X G G mixture = + ( 1 1 2 2 ) + ∆ mixture

where X1 and X2 are the mole fractions of two components 

in a binary mixture. The ∆Gmixture term contains an ideal 

component that depends on X1 and X2 and a non-ideal 

component dependent on X1, X2, and activity coefficients 

in the solid phase (γ1 and γ2):

∆G RT X X X X

RT X X

mixture = [ + ]+

[ + ]

1 1 2 2

1 1 2 2

ln ln

ln ln γ γ

For an ideal solid solution, γ1 = γ2 = 1, so that 

RT [X1 ln γ1+X2 ln γ2] = 0. Ideal mixing may be approached 

where the amount of substitution is very low (a dilute solid 

solution) or where the mixing cations are closely matched 

in size and charge. In this case, the ∆Gmixture function is a 

symmetrical parabola having a minimum at X1 = X2 = 0.5. 

Hence as a general rule, the free energy of binary mixtures 

is less than that of the pure, end-member components. It 

follows that the solubility of an ion can be lowered in a 

mixed ionic compound relative to the solubility of the pure 

compound.

Remobilization of a coprecipitated contaminant will be governed by the overall stability of the host precipitate, which 

may be controlled by ground water parameters such as pH 

and/or redox state. In most cases, the identification of a 

coprecipitation process cannot be made with a single line 

of evidence. Observations of decreased contaminant concentrations concurrent with decreases in the concentrations 

of major ground-water constituents such as Ca, Fe, or dissolved sulfide may be indicative of a coprecipitation event. 

It is recommended that this evidence be supplemented with 

solid phase characterization approaches, such as chemical 

extractions or microanalytical techniques, to confirm that 

coprecipitation is an attenuation mechanism. 

IIB.2.3 Surface Precipitation

Surface precipitation may result when adsorption leads 

to high sorbate coverage at the mineral-water interface. 

Surface precipitation can be thought of as an intermediate 

stage between surface complexation and bulk precipitation 

of the sorbing ion in solution (Farley et al., 1985; Ford et 

al., 2001). At low concentrations of the sorbing metal at 

the mineral surface, surface complexation is the dominant 

process. As the concentration of the sorbate increases, 

the surface complexation concentration increases to the 

point where nucleation and growth of a surface precipitate 

occurs. Surface precipitation can be viewed as a special 
36

case of coprecipitation where the mineral interface is a 

mixing zone for ions of surface precipitate and those of the 

underlying substrate. It is generally believed that surface 

precipitation can occur from solutions that would appear to 

be undersaturated relative to precipitate formation based 

on considering solution saturation indices. The reason for 

this may be due in part to different equilibrium constants 

for surface precipitation versus precipitation from solution, 

or may be related to the way the mineral-water interface is 

modeled (Sverjensky, 2003). For example, the dielectric 

constant of water and therefore activity coefficients in bulk 

solution may be different from activity coefficients near 

a mineral surface. Again, the identification of a surface 

precipitation process cannot be made only with solution 

data. Solid phase characterization data, such as chemical extractions or microanalytical techniques, are needed 

to confirm that surface precipitation plays a role as an attenuation mechanism.

IIB.2.4 Mineral Transformation

In many cases the solids that precipitate in near surface 

environments are not the most thermodynamically stable 

phases. For example, hydrous ferric oxide, ferrihydrite, 

is metastable relative to the iron oxyhydroxide goethite. 

The preponderance of metastability in near surface environments is a consequence of the slowness of chemical 

reactions at temperatures typical of surficial environments. 

Kinetics, therefore, play an integral role in ground-water and 

soil geochemistry. Mineral transformation is one example 

of how metastable precipitates evolve toward more stable 

mineral phases within an aquifer. Ultimately, contaminants 

that are initially adsorbed onto or coprecipitated with these 

metastable precipitates are likely to become more resistant 

to remobilization if they are incorporated into the more 

stable transformation product.

The Ostwald Step Rule is often obeyed in low-temperature 

mineral formation. Precipitation of less stable and 

more soluble phases is followed by transformation to 

progressively more stable and less soluble phases. This 

behavior stems from the preferential formation of materials 

with fast precipitation kinetics over nucleation and growth 

of phases with slow kinetics (Stumm, 1992). Differences in 

precipitation kinetics are often tied to structural complexity 

of the precipitating mineral. Relatively simple structures are 

able to form rapidly whereas ordered structures, although 

more stable, require longer time periods to develop. 

Precursor phases are usually poorly crystalline and they 

may be chemically dissimilar to the final stable mineral. 

Examples that follow the Ostwald Step Rule include the 

precipitation of ferrihydrite and transformation to more stable 

iron oxyhydroxides (goethite) and iron oxides (hematite), 

the precipitation of mackinawite and transformation to 

pyrite, and the precipitation of amorphous aluminosilicates 

such as allophane and transformation to halloysite and 

kaolinite. Transformation pathways result from solution 

mediated processes or solid-phase transitions (Steefel and 

van Cappellen, 1990).

The iron monosulfide-to-iron disulfide transformation has 

been widely studied in the laboratory and in the field. In this 

example of a mineral transformation process, mackinawite 

(Fe1+xS) precipitates as concentrations of dissolved sulfide 

and ferrous iron accumulate in pore water. It has been determined that in sulfate-reducing environments, pore water 

concentrations of ferrous iron and sulfide are controlled 

by the solubility of mackinawite. Mackinawite, however, is 

metastable with respect to the iron disulfides, pyrite and 

marcasite. The rate of transformation from mackinawite 

to pyrite or marcasite depends on pH and redox conditions. Metals that coprecipitate with mackinawite are likely 

incorporated into pyrite, which is more stable over a wide 

pH range and in anoxic conditions. The rate at which this 

transformation occurs will be governed by chemical conditions including the coprecipitation or adsorption of contaminants and other dissolved constituents from solution. Site 

characterization aspects relating to mineral transformation 

processes as an immobilization mechanism will involve 

determining the spatial concentration distribution of precursor phases and their more stable transformation products 

along with contaminant associations using mineralogical 

and wet chemical characterization tools.

IIB.3 Implications for Natural Attenuation 

Assessment

The sorption processes discussed in the preceding paragraphs may act in isolation or together to arrest contaminant 

migration within the aquifer. Factors that dictate which process is likely to dominate contaminant attenuation include 

chemical properties of the contaminant, chemical characteristics of the ground water, and properties of the aquifer 

solids. Due to the complexities of directly identifying the 

immobilized form of the contaminant, it is likely that multiple 

lines of evidence will be needed to adequately discern the 

controlling attenuation reaction. These lines of evidence will 

include the evaluation of patterns in ground-water chemistry 

that point to potential precipitation or coprecipitation reactions, evaluation of aquifer solids to determine patterns in 

contaminant and solid component associations, and the 

use of chemical speciation or reaction models to assess 

if ground-water and aquifer solid characteristics are consistent with observed contaminant attenuation. Additional 

perspective on possible sorption processes for specific contaminants is provided in the contaminant-specific chapters 

included in Volume 2 and 3 of this document. 

IIC. Microbial Impacts on Inorganic 

Contaminant Attenuation

The chemical characteristics of ground water and properties 

of the aquifer mineral components are, in part, influenced 

by microbial reactions. Microbial activity within the aquifer 

may also play a more direct role in controlling contaminant 

speciation and migration. The influence of microbial reactions may be more pronounced in contaminant plumes that 

also contain degradable organic contaminants such as 

hydrocarbons or chlorinated solvents. In these instances, 

the plume geochemistry may differ significantly from that 

observed in ambient ground water at a site. If microbial 

reactions play a significant role in contaminant attenuation, it 

may be necessary to gather information on the degree that 
37

these reactions control the concentrations and distributions 

of reactants that participate in contaminant attenuation, 

as well as the capacity of the aquifer to sustain microbial 

activity within the plume. In order to provide perspective 

on the potential influence of the subsurface microbiology, 

the following discussion provides an overview of general 

characteristics of subsurface microbiology, the influence of 

microbial activity on the redox state within the plume, and 

the types of contaminant attenuation reactions that may be 

directly mediated by microorganisms.

IIC.1 Characteristics of Aquifer Microbiology

Microbial reactions can alter the geochemical structure 

of the subsurface (Newman and Banfield, 2002), and it 

is becoming increasingly important to characterize these 

processes in terms of the natural attenuation of inorganic 

contaminants. Microorganisms inhabiting the subsurface 

environment exhibit a remarkable array of metabolic capabilities enabling them to use organic or inorganic matter as 

energy sources and propagate under aerobic or anaerobic 

conditions. A large part of a microorganism’s metabolism 

is devoted to the generation of adenosine triphosphate 

(ATP), an “energy currency” that is used by the microorganism for cell synthesis. Microorganisms derive energy 

through the oxidation of organic compounds or chemicallyreduced inorganic ions and compounds. The electrons or 

hydrogen atoms resulting from oxidation are transferred 

in most microorganisms by an electron transport chain to 

an electron acceptor which, in the case of aerobic respiration, is oxygen. Some microorganisms are capable of 

anaerobic respiration, whereby the electron acceptor is not 

O2 but chemically-reducible inorganic compounds such as 

NO3

-

, SO4

2-, CO2

, Fe3+, or iron/manganese oxyhydroxides. 

A third type of metabolism, called “fermentation,” involves 

intra-molecular oxidation/reduction without an externally 

supplied terminal electron acceptor. Fermentation reactions are not considered here because they do not involve 

metal transformations.

The metabolic diversity manifested by microorganisms is 

key to their being agents of geochemical change in the 

subsurface (Ehrlich, 1995). Subsurface biota are usually 

categorized as either heterothropic or chemolithotrophic 

according to the following criteria:

• Heterotrophic microorganisms - Heterotrophs include 

bacteria (single- or multi-celled organisms lacking 

internal membrane structures), fungi (mycelial or 

single-celled organisms possessing a cell wall but 

no photosynthetic capability), protozoa (unicellular, 

microscopic animals such as amoebae), and archaebacteria (a newly discovered group of organisms possessing a unique cell envelope, membrane structure 

and ribosomal RNA, distinguishing them from all other 

microorganisms). Heterotrophs derive energy from the 

oxidation of organic compounds and obtain most of their 

carbon from these compounds. Some heterotrophs 

respire aerobically while others respire anaerobically, 

requiring the availability of oxygen or other reducible 

compounds to serve as electron acceptors. 

• Chemolithotrophic microorganisms - Some genera 

of bacteria and archaebacteria are chemolithotrophs. 

These microorganisms derive energy from the oxidation of inorganic compounds (for example, Fe2+, HS-

, 

and H2) and obtain their carbon as CO3

2-, HCO3-

 or 

CO2. Some chemolithotrophs are aerobic using O2 as 

an electron acceptor while others respire anaerobically using chemically reducible inorganic matter as 

an electron acceptor (Newman, 2001). An example of 

this type of metabolism is carried out by methanogens, 

which oxidize H2 using CO2 as an electron acceptor to 

form CH4. Reactions carried out by chemolithotrophic 

microorganisms must derive sufficient energy from 

the oxidation/reduction reactions to fix carbon (that is, 

reduce CO3

2-, HCO3-

 or CO2 to organic carbon) and 

phosphorylate ADP.

In the aquifer below the vadose zone the populations and 

numbers of microorganisms vary significantly due to limitations to available oxygen. If the aquifer is shallow and is 

recharged through cracks and fissures, the water may be 

nutrient-rich and oxygenated resulting in an abundance of 

different microorganisms whose numbers can reach 106

per gram. Lower population numbers would be expected 

in aquifers where the supply of dissolved oxygen from surface recharge is limited. Deep aquifers are generally more 

depleted with respect to nutrients (unless contaminated) 

and oxygen causing the numbers and diversity of microorganisms to decline and often be limited to bacteria and 

archaebacteria. This finding has been supported, during 

the last 15–20 years, by carefully controlled drilling studies 

undertaken to evaluate the microbiology of deep, impermeable rock strata (Ghiorse, 1997; Kerr, 2002).

IIC.2 Microbial Controls on Subsurface Redox 

State

Since microorganisms chemically transform aquifer constituents such as dissolved oxygen, iron [aqueous and solid 

forms of Fe(III) and Fe(II)], and sulfur [aqueous species such 

as sulfate and dissolved sulfide], their metabolic reactions 

may exert significant influence on the redox geochemistry 

within the plume. As previously discussed, redox conditions within the aquifer may govern precipitation-dissolution reactions that control contaminant precipitation or 

coprecipitation, as well as the types and concentration of 

aquifer solids that may serve as adsorbents. From this 

perspective, some knowledge about the microbial populations as a function of space and time within the plume may 

be necessary to understand the existing redox status of the 

aquifer and make projections about how it may evolve. For 

heterotrophic microorganisms, the electrons or reducing 

equivalents (hydrogen or electron-transferring molecules) 

produced during degradation of organic compounds must 

be transferred to a terminal electron acceptor (TEA). Observations of microbial systems have led to the development 

of a classification system that groups microorganisms into 

three categories according to predominant TEAs:

• Aerobic bacteria — Bacteria which can only utilize molecular oxygen as a TEA. Without molecular oxygen, 

these bacteria are not capable of degradation. 
38

• Facultative aerobes/anaerobes — Bacteria which can 

utilize molecular oxygen or when oxygen concentrations 

are low or nonexistent, may switch to nitrate, manganese oxides or iron oxides as electron acceptors.

• Anaerobes — Bacteria which cannot utilize oxygen 

as an electron acceptor and for which oxygen is toxic. 

Though members may utilize nitrate or other electron 

acceptors, it can be said that they generally utilize 

sulfate or carbon dioxide as electron acceptors.

In any environment in which microbial activity occurs, there 

is a progression from aerobic to anaerobic conditions (ultimately methanogenic) with an associated change in the 

redox status of the system. There is a definite sequence 

of electron acceptors used in this progression through 

distinctly different redox states. The rate, type of active 

microbial population, and level of activity under each of 

these environments are controlled by several factors. 

These include the concentration of the electron acceptors, substrates which can be utilized by the bacteria, and 

specific microbial populations leading to the progression 

of an aquifer from aerobic to methanogenic conditions 

(Salanitro, 1993). This results in a loss of organic carbon 

and various electron acceptors from the system as well as 

a progression in the types and physiological activity of the 

indigenous bacteria.

If microbial activity is high, the aquifer environment would 

be expected to progress rapidly through these conditions. 

The following scenario outlines a general sequence of 

events in which aerobic metabolism of preferential carbon 

sources would occur first. The carbon source may be from 

organic contaminant sources or other more readily degradable carbon which has entered the system previously or 

simultaneously with the contamination event.

• Oxygen-Reducing to Nitrate-Reducing Conditions 

– Once available oxygen is consumed, active aerobic 

populations begin to shift to nitrate respiration. Denitrification will continue until available nitrate is depleted, 

or usable carbon sources become limiting. 

• Nitrate-Reducing to Manganese-Reducing Conditions 

– Once nitrate is depleted, populations which reduce 

manganese may dominate. Bacterial metabolism of 

substrates utilized by manganese-reducing populations 

will continue until the concentration of manganese oxide 

becomes limiting. 

• Manganese-Reducing to Iron-Reducing Conditions 

– When manganese oxide becomes limiting, iron 

reduction becomes the predominant reaction mechanism. Available evidence suggests that iron reduction 

does not occur until all Mn(IV) oxides are depleted. In 

addition, bacterial Mn(IV) respiration appears to be 

restricted to areas where sulfate is nearly or completely 

absent. 

• Iron-Reducing to Sulfate-Reducing Conditions – Iron 

reduction continues until substrate or carbon limitations allow sulfate-reducing bacteria to become active. 

Sulfate-reducing bacteria then dominate until usable 

carbon or sulfate limitations impede their activity. 

• Sulfate-Reducing to Methanogenic Conditions – Once 

usable carbon or sulfate limitations occur, methanogenic bacteria are able to dominate.

Development of a general knowledge of the redox status 

of the aquifer throughout the plume is important relative to 

understanding the processes contaminant attenuation (or 

lack thereof) within the plume, as well as changes in the 

capacity for or stability of contaminant attenuation with the 

return to ambient conditions. There are several approaches 

to evaluating the redox state of the aquifer. One common 

approach is to monitor the oxidation-reduction potential in 

ground water using a platinum electrode with recalculation 

of the measured electrode potential as Eh. However, due 

to the lack of internal redox equilibrium in natural systems 

(Morris and Stumm, 1967) and limitations to electrode potential measurements in a complex matrix such as ground 

water, this measurement should be supplemented with other 

site-specific determinations. One supplemental approach to 

assessing redox status is the determination of concentrations of specific redox sensitive species such as oxygen, 

Fe(II), hydrogen sulfide, or methane as qualitative guides 

to the redox status of ground water (Lindberg and Runnells, 

1984; USEPA, 2002). These measurements provide a 

means to generally characterize the ground water as oxic, 

suboxic, or anoxic and may provide more direct evidence of 

the occurrence of iron-reducing, sulfate-reducing or methanogenic conditions. Another approach that may be used 

to indicate the terminal electron acceptor process (TEAP) 

which dominates within the plume is measurement of the 

hydrogen (H2

) concentration in ground water (Lovely and 

Goodwin, 1988). Hydrogen concentration for the various 

TEAPs are shown in Table 2.4 (Chapelle et al., 1995).

Table 2.4 Range of hydrogen concentrations for a 

given terminal electron-accepting process 

that can be used for classification of the 

redox status within the contaminant plume.

Terminal Electron 

Accepting Process

Hydrogen (H2) 

Concentration 

(nanomoles per liter)

Denitrification <0.1

Iron (III) Reduction 0.2 to 0.8

Sulfate Reduction 1 to 4

Methanogenesis 5 to 20

The focus of this discussion has been on the influence of 

microbial activity on the redox status of ground water, which 

governs the types of aqueous and solid phase reactants that 

may be involved in contaminant attenuation. The following 

discussion elaborates on the influence of microbial reactions 

on the chemical speciation of inorganic contaminants.
39

IIC.3 Impacts on Contaminant Speciation and 

Attenuation

Microbe/metal interactions result in immobilization, mobilization, or transformation by extracellular precipitation reactions, intracellular accumulation, oxidation and reduction 

reactions, methylation and demethylation, and extracelluar 

binding and complexation. Many of these processes play 

an important role in achieving natural attenuation through 

immobilization or degradation. Many transformations of 

metals and metalloids are carried out by subsurface microorganisms for the purpose of obtaining energy for growth 

and reproduction; however, these transformations may not 

result in decreasing the mobility or toxicity of the metal. 

Some microbial transformations are carried out specifically 

to detoxify their environment, suggesting that microbes have 

an enormous capacity to sustain life. Therefore, the role of 

microbes in the subsurface is multifaceted. The principal 

mechanisms employed by microorganisms to transform 

metals and metalloids in the subsurface environment are: 

• Oxidation and reduction reactions resulting from microbial respiratory activities (Lovely, 1993; Ahmann et 

al., 1994);

•	 Biosorption by cell walls, cell membranes and exopolymers and intracellular bioaccumulation (Beveridge, 

1989); and 

•	 Methylation and demethylation (Oremland et al., 1991; 

Bentley et al., 2002).

These microbial mechanisms are described with emphasis 

on reactions involved in natural attenuation of metals and 

radionuclides.

IIC.3.1Contaminant Oxidation-reduction 

Reactions

Oxidation-reduction (redox) processes are chemical reactions that involve a transfer of electrons between reactants 

and products and consequently a change in the oxidation 

state of elements that can either be oxidized to higher valence states or reduced to lower valence states. Changes 

in the valance state of a particular element can result in 

the following:

• Change in speciation resulting in lower or higher toxicity

• Change in speciation resulting in lower element solubility (immobilization) or increased solubility (mobilization)

• Change in speciation that impacts adsorption/desorption behavior

The contaminants for which redox-mediated processes 

may lead to attenuation include elements that can occur 

in multiple oxidation states under environmentally relevant conditions. Thus, a transition metal such as Cr is 

redox-active since two stable oxidation states (VI and III) 

are readily observed in aqueous solution, whereas Ni is 

not redox-active as only one oxidation state (II) occurs in 

aqueous solutions at ambient conditions. (More detailed 

discussions on the redox activity of individual contaminants 

are provided in Volume 2.)

 Oxidation-reduction (redox) reactions, which can be carried 

out by both heterotrophic and chemolithotrophic microorganisms, are either energy-generating or enzymaticallycatalyzed without coupling to energy generation. Microbially catalyzed, energy generating redox reactions tend to 

be rapid and the metal transformations so extensive that 

the geological influence is often substantial. One example 

that results in the attenuation of metals is the reduction of 

soluble SO4

2- by Desulfovibrio and Desulfotomaculum species involving oxidation of organic matter or H2

 for energy 

and use of sulfate as a terminal electron acceptor (Lovely, 

1993). This redox reaction yields dissolved sulfide that 

can chemically react with a number of soluble, divalent 

metals producing insoluble metal sulfides. In some cases, 

the microbially-generated sulfide may chemically reduce 

an inorganic contaminant. For example, dissolved sulfide 

may result in reduction of arsenate to arsenite. Examples 

of enzymatically-catalyzed redox reactions that immobilize 

metals for attenuation in the subsurface are the reduction 

of mobile Cr(VI) to the less toxic and less mobile Cr(III) 

(Palmer and Wittbrodt, 1991) and reduction by some 

microorganisms of Se(VI) or Se(IV) to insoluble Se(0) 

(Ehrlich, 1995).

IIC.3.2Biosorption and Intracellular 

Bioaccumulation

Biosorption, a form of contaminant adsorption, is the binding 

of inorganic ions to the outer surface of a microorganism by 

ion exchange, van der Waals attractions, or chemical reactions between the metal ion and cell wall, cell membrane, 

excreted metabolites, or exopolymers. Biosorption is not 

dependent on metabolism by the microbial cell. Intracellular bioaccumulation, or absorption, is the transport of 

metals across the cell wall and membrane and subsequent 

retention of the metals within the cell. Bioaccumlation is 

metabolism dependent because the transport process 

requires energy.

The outer surfaces of microorganisms are chemically 

complex with each group of organisms possessing differing 

cell wall, membrane, and expolymeric structures. These 

structures are composed of macromolecules having 

functional groups including carboxylates, amines, imidazole, 

phosphate, sulfydryl, sulfate, hydroxyl and others. Usually 

the microbes exhibit a net negative charge owing to the 

abundance of carboxylate and phosphate groups, however, 

positively charged amines and imidazole functional groups 

impart polyfunctional characteristics to the cell surface. 

Most inorganic contaminant binding occurs after initial 

complexation and neutralization of the chemically active 

site. This stoichiometric interaction between contaminant 

ions and active sites within the wall acts as a nucleation 

site for the deposition of additional ions from solution in a 

manner similar to surface precipitation onto aquifer minerals. 

The deposition product then grows in size within the 

intermolecular spaces of the wall fabric until it is physically 
40

constrained by the polymeric meshwork of the wall. 

Therefore, the amount of contaminants bound to the outer 

surfaces of organisms can be considerably greater than that 

accounted for by the number of charges available because 

binding occurs not only by ion exchange type reactions 

and covalent bonding to charged functional groups, but by 

precipitation and reduction reactions (Beveridge, 1989). 

As with mineral surfaces in the aquifer solids, pH plays a 

critical role in the binding of inorganic contaminants. At 

low pH values cationic metal binding is diminished because 

negative charges on functional groups become protonated, 

reducing the negative charge density of the organism’s surface. Although the quantitative role of extracellular binding 

in controlling the concentration of inorganic contaminants 

in the subsurface is not well defined, it does play a larger 

part in metal mobility than intracellular metal accumulation. 

The relative contribution of microbial binding processes to 

the overall immobilization of the contaminant within the 

plume is largely unknown. However, a simple comparison 

of the relative mass of microbes versus reactive minerals 

in aquifer solids would suggest that microbial binding plays 

a minor role.

IIC.3.3 Methylation and Demethylation

Microbial interactions with some inorganic contaminants 

considered in this document result in the transfer of methyl 

groups onto these elements by a process known as methylation. This process may result in either the increased 

toxicity or mobility of some inorganic contaminants. Biomethylation of metal(loid)s takes place when suitable organisms are present under anaerobic conditions with high 

concentration of available metal(loid)s and methyl donors. 

These conditions may exist in contaminated ground water 

(e.g., derived from landfill leachate). An example of this 

type of process is the methylation of inorganic selenium 

to dimethylselenide, a volatile product. Methylation of 

selenium decreases rather than increases its toxicity and 

probably represents a detoxification reaction that occurs 

readily in aerobic soils and sediments. Arsenate can also 

be methylated by some bacteria and filamentous fungi. 

Biosynthesis of organic compounds containing arsenic such 

as arsenocholine, arsenobetaine, and arsenosugars also 

occurs among various bacteria, algae, and many higher 

organisms (Andreae and Klumpp, 1979).

IIC.4 Implications for Natural Attenuation 

Assessment

Based on the previous discussion, it is apparent that the 

subsurface microbial community could influence inorganic 

contaminant transport. For many of the contaminants discussed in this document, microbial reactions will primarily 

exert an indirect influence on contaminant attenuation by 

governing the types, distribution, and concentrations of 

aqueous and solid phase reactants. However, microbial 

activity may constitute a more direct impact on contaminant 

speciation, including changes in the redox state of contaminants such as arsenic or selenium. From this perspective, 

it is recommended that the site characterization effort incorporate analysis of the extent to which microbial processes 

may govern ground-water chemistry within the plume. This 

is more critical at sites where the co-occurrence of readily 

degradable organic contaminants result in ground-water 

conditions significantly different than those observed in 

ambient ground water. For these situations, information on 

the mechanism, rate, capacity, and stability of contaminant 

attenuation may need to be evaluated within the context of 

the behavior of the organic component of the plume.

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Farley, K.J., D.A. Dzombak, and F.M.M. Morel. A surface 

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Ford, R.G. Rates of hydrous ferric oxide crystallization and 

the influence on coprecipitated arsenate. Environmental

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metal sorption/precipitation mechanisms on soil mineral 

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Gschwend, P.M. and M.D. Reynolds. Monodisperse ferrous 

phosphate colloids in an anoxic groundwater plume. 

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Lindberg, R.D. and D.D. Runnells. Ground water redox 

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Lowers, H.A., G.N. Breit, A.L. Foster, J. Whitney, J. Yount, 

Md.N. Uddin, and Ad.A. Muneem. Arsenic incorporation into authigenic pyrite, Bengal Basin sediment, 

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43

Section III

Site Characterization to Support Evaluation of MNA

Richard T. Wilkin, Steven Acree, Steve Mangion, Patrick V. Brady, James E. Amonette, Paul 

M. Bertsch, Douglas B. Kent, Robert W. Puls, Ann Azadpour-Keeley, Robert G. Ford 

In determining whether MNA is applicable to a site, characterization data are collected to define site hydrogeology 

and the reaction(s) that result in contaminant attenuation. Characterization of attenuation reactions will involve 

identification of the reactants that control contaminant 

immobilization or degradation. The reactants may be of 

aqueous and/or solids forms, with solid reactants potentially 

consisting of aquifer solids or microbes. Once the specific 

reaction mechanism is identified, site characterization will 

then proceed to determination of the capacity of groundwater conditions to sustain the attenuation reaction and 

assessment of the stability of immobilized contaminants 

to resist remobilization due to changes in ground-water 

chemistry. The types of samples collected from within 

the plume boundary will include ground water and aquifer 

solids. Analyses of these samples will include those more 

commonly employed to understand the overall geochemical 

context of subsurface chemistry, as well as more specialized 

analyses used to identify the chemical speciation of the 

contaminant in dissolved or solid form. Simple observations of the distribution of the contaminant between solid 

and liquid forms will be insufficient to identify the specific 

reaction process. A mechanistic understanding of the reaction is needed in order to assess the overall performance 

characteristics of the attenuation process. This section 

will provide an overview of the analytical techniques that 

are used to provide evidence for processes and chemical 

speciation in the subsurface.

IIIA Site Hydrogeology

Hydrogeology is the foundation of the conceptual model 

for natural attenuation (National Research Council, 2000). 

Three-dimensional characterization of the ground-water 

flow field and the changes that occur through time are 

crucial to understanding the transport and, ultimately, the 

fate of contaminants. Discussions of hydrogeologic site 

characterization in this document draw freely from USEPA 

(2003) and are largely limited to evaluations performed in 

saturated porous-media settings, such as unconsolidated 

aquifer materials. Although similar concepts are sometimes 

applicable in other geologic settings, such as fractured rock 

formations, ground water often moves primarily through 

discrete pathways at these sites. In such highly anisotropic 

settings, the direction of ground-water flow and contaminant 

transport often cannot be determined within a sufficient 

degree of certainty to support assessments of natural attenuation rates or processes. Assessment of contaminant 

transport and fate in these and other highly heterogeneous 

settings is an area of continuing research and is beyond the 

scope of this discussion. In similar fashion, hydrogeologic 

characterization of vadose zone materials is not specifically 

addressed in this section. Flow conditions are often more 

complex in the vadose zone than in similar saturated media. 

However, characterization of the hydrogeologic properties 

of the vadose zone, as well as the saturated zone, will be 

necessary at some sites. A general discussion of vadose 

zone characterization for contaminant transport assessments is found in USEPA (1991; 1993a; 1993b).

IIIA.1 Characterization Objectives

Much of the spatial variability in observed contaminant 

concentrations is the result of geologic heterogeneity. Differences in the geologic properties of aquifer materials from 

the micro-scale (e.g., grain-size distribution) to the field 

scale (e.g., stratigraphy) result in differences in hydraulic 

conductivity and the directions and rates of ground-water 

flow. Three-dimensional hydraulic gradients that occur due 

to the site-specific characteristics of ground-water recharge 

and discharge are also dominant controls on flow. Together, 

the differences in the hydraulic conductivity of different 

aquifer materials, hydraulic gradients, and the changes in 

gradients that often occur in response to natural or anthropogenic changes in ground-water discharge and recharge 

control the directions and rates of ground-water flow.

Detailed hydrogeologic characterization is essential for 

evaluating natural attenuation processes and rates, as 

well as for specification of the locations and frequencies 

for performance monitoring. The first step in characterizing contaminant transport, and, therefore the processes 

that may result in attenuation, is the determination of the 

path or paths of ground-water transport along which the 

contaminant will migrate. At a minimum, the hydrogeologic 

database should be sufficient to:

• Define geologic and hydrologic controls on the groundwater flow field (e.g., transmissive units, preferential 

pathways, and barriers to flow);

• Quantify flow rates and directions and their spatial and 

temporal variations;

• Support characterization of contaminant sources, the 

distribution of contaminants of concern, and the effects 

of dominant natural attenuation processes; and

• Identify possible receptors.
44

The scale and intensity of this characterization is determined by site-specific conditions (e.g., variability in geology 

and hydrology and the acceptable level of uncertainty in 

the outcome of the evaluations). 

IIIA.2 Geologic Characterization

IIIA.2.1 Saturated Porous Media

The key geologic factors that control ground-water flow at 

sites where secondary porosity features such as fractures, 

including faults, joints, and partings along bedding planes, 

are not present include site stratigraphy and lithologic 

characteristics. Stratigraphy is the science concerned with 

the form, composition, thickness, areal extent, sequence, 

and correlation of both consolidated and unconsolidated 

geologic materials. It defines the framework of the groundwater flow field and constrains the pathways for contaminant 

migration (USEPA, 1991). Lithology is the description of 

the physical and mineralogic characteristics of geologic 

materials. Physical characteristics that control flow include 

grain-size distribution, grain shape, and packing. As discussed in other sections, mineral composition and coatings 

affect inorganic constituent distribution through a variety of 

chemical reactions including cation exchange, substitution, 

precipitation, dissolution, acid-base reactions, complexation 

and redox reactions (USEPA, 1991). Grain-size distribution, 

grain shape, and packing (i.e., arrangement) of grains influence the hydraulic conductivity, total porosity, and effective 

porosity of the geologic materials. 

At many sites, evaluation of sedimentary depositional 

environments is an especially useful framework for the 

understanding of site stratigraphy and the distribution of 

lithologic controls on ground-water flow. Sedimentary facies 

(i.e., sedimentary bodies that are internally similar in characteristics) determine the three-dimensional geometries, 

connectivity, and heterogeneity of aquifer units and barriers 

to flow (Galloway and Sharp, 1998) in many porous-media 

settings. Ground water can move preferentially through 

coarser-grained materials, resulting in varying degrees of 

heterogeneity in flow patterns and contaminant transport. 

Interconnected facies of relatively coarse-grained materials may provide preferential pathways for contaminant 

migration. An example of a naturally occurring preferential 

pathway would be a deposit of interconnected sands of high 

hydraulic conductivity bounded by deposits of finer-grained 

materials. In similar fashion, anthropogenic features such 

as buried utility corridors and heterogeneous fill materials 

may also result in the formation of preferential pathways for 

ground-water flow or impediments to flow. However, even 

within a given preferential pathway, ground water may move 

in sinuous paths due, in part, to small-scale differences 

in hydraulic conductivity of the materials or to temporally 

variable, three-dimensional hydraulic gradients.

Interconnected transmissive materials may be separate and 

distinct pathways for contaminant movement. For example, 

the degree of hydrologic connection between different sedimentary facies depicted in Figure 3.1 is small. Monitoring 

points in different facies may appear to be similar and may 

be hydraulically down gradient of one another without significant ground-water flow and contaminant transport from 

one unit to the next. This is explicitly illustrated in Figure 

3.1 where two wells are screened at similar depth below the 

water table along the pathway of ground-water transport. 

The two screens are placed within two different facies that 

likely possess different magnitudes of hydraulic conductivity 

(i.e., Well 1 screened in a higher conductivity material in 

the “medium to coarse-grained sand” vs. Well 2 screened 

in a lower conductivity in the “fine-grained silty sand”). The 

difference in hydraulic conductivity between these two facies may prevent a direct line of transport from Well 1 to 

Well 2. In such cases, apparent contaminant attenuation 

between monitoring points (i.e., due to lower concentration 

observed in Well 2) may be an artifact of sample location 

and not representative of actual conditions. Inferences 

about natural attenuation based on apparent decreases in 

contaminant concentration in the down gradient direction 

are likely to be incorrect in these situations unless groundwater flow paths are determined and monitored. This level 

of characterization is often difficult to accomplish using 

small numbers of monitoring points. 

Three-dimensional characterization will be needed to evaluate and predict the effects of natural attenuation processes 

(e.g., advection, dispersion, diffusion, and sorption processes) on contaminants at many sites. Data required to 

construct the site geologic framework include stratigraphic 

and lithologic data obtained from geologic cores and supplemented with information from surface and, particularly, 

borehole geophysical methods. Innovative characterization 

technologies, such as the cone penetrometer and geologic 

sampling using direct-push methods, offer the potential for 

cost-effectively evaluating the geologic controls on groundwater flow and their variability in greater detail than previously possible without obtaining continuous cores during 

traditional drilling. Regardless of the chosen technology, 

the choice of sampling locations needed as a basis for 

evaluations of natural attenuation processes depends 

on factors such as the site-specific geology, contaminant 

characteristics, degree of physical, chemical, and biological 

heterogeneity, and the locations of points of compliance 

and or critical down gradient exposure points. 

IIIA.2.2 Saturated Fractured Media 

In rock or consolidated materials, features such as fractures often control ground-water flow. The primary factors 

affecting flow through fractured media are the density and 

orientations of fractures, the effective aperture width of the 

fractures, and properties of the rock matrix (Schmelling and 

Ross, 1989). An adequate characterization of a fractured 

rock system would generally include information on fracture 

orientation (i.e., strike and dip), aperture widths, fracture locations, and interconnection; hydraulic head throughout the 

contaminated volume; the distribution of rock porosity and 

permeability; sources of ground water and contaminants; 

contaminant distribution; and chemical interactions with 

the rock matrix. Chemical weathering of the rock and the 

subsequent mineralogic changes that occur along fractures 

can also be especially important controls on the fate and 

transport of inorganic contaminants.
45

Many important features, such as fracture patterns, can 

often be determined by standard surficial geologic mapping 

techniques. Much additional data concerning subsurface 

conditions can also be obtained using downhole and surface 

geophysical methods. However, sufficient characterization 

to support natural attenuation process evaluations with the 

level of certainty needed to satisfy data quality objectives 

will be neither technically nor economically feasible at the 

scale of many sites. 

IIIA.3 Hydrologic Characterization

Knowledge of the geologic framework provides much information needed to define the features that control groundwater flow. However, data regarding the hydraulic properties 

(e.g., hydraulic conductivities and effective porosities) of the 

aquifer materials and their distribution are needed to allow 

quantitative estimates of such parameters as ground-water 

flow rates, contaminant fluxes, and contaminant attenuation rates. Much of the necessary data regarding hydraulic 

properties will generally be acquired through hydraulic tests 

performed in the field (e.g., pumping tests, slug tests, packer 

tests, and tracer tests). Additional information regarding 

geologic and hydrologic site characterization concepts and 

techniques may be obtained from a variety of sources (e.g., 

Butler, 1998; Kruseman and de Ridder, 1989; USEPA, 1991; 

USEPA, 1993a; USEPA, 1998).

Hydrologic characterization of the ground-water flow 

system also requires an understanding of natural and 

anthropogenic sources for recharge, the characteristics of 

discharge, and the hydraulic gradients that are the result. 

Recharge sources include precipitation, surface-water 

bodies, irrigation, and losses from potable water distribution 

systems. Important locations of ground-water discharge 

include water production wells, springs, wetlands, and 

surface-water bodies. The locations and rates of groundwater recharge and discharge are important factors in 

determining site-specific hydraulic gradients and, therefore, 

the directions and rates of flow. Hydraulic gradients may be 

three dimensional in nature with strong vertical as well as 

horizontal components in many cases. This often results 

in a dynamic, three-dimensional flow system.

Figure 3.1 Geologic block diagram and cross section depicting a stream environment. In this setting, numerous

facies of channel materials are surrounded by finer-grained materials of lower hydraulic conductivity

(modified from USEPA, 2003). Two monitoring wells, labeled [1] and [2], are shown in red. A magnified

portion of the aquifer cross-section is shown to the right to clarify that the well screens are placed within

two different facies possessing differing hydraulic conductivites.
46

Temporal variations in either natural or anthropogenic 

recharge or discharge may result in fluctuations in both the 

horizontal and vertical components of hydraulic gradients, 

the directions/rates of contaminant migration, contaminant 

loading to ground water, and, potentially, redox conditions. 

These fluctuations may lead to changes in the geometry 

of the plume that should be considered during evaluations 

of natural attenuation and monitoring plan development. 

For example, seasonal changes in precipitation may result 

in non-uniform changes in water table elevations due to 

differences in recharge related to topography, different soil 

types, or land uses. This may result in seasonal changes 

in hydraulic gradients or discharge locations. Rapid and 

sustained changes in ground-water flow velocity are 

common in flood plains, particularly near large rivers that 

have major changes in the river stage. In some situations, 

changes in contaminant loading to ground water may also 

occur due to increased recharge through contaminated 

vadose zone materials or elevation of the water table into 

these materials. Longer-term patterns associated with 

sequences of unusually wet or dry years may also influence 

ground-water velocity over correspondingly longer time 

periods. 

Anthropogenic influences on site hydrology, such as 

changes in ground-water withdrawal or irrigation rates and 

patterns, may have similar effects on plume behavior but 

occur on frequencies other than those corresponding to 

precipitation patterns. Irrigation or municipal water supply wells that pump intermittently can affect ground-water 

flow patterns in a complex manner that may be difficult to 

assess. In addition, land use changes that alter patterns 

of recharge, discharge, or withdrawal may be important 

sources of variability in the ground-water flow field that 

should be routinely considered during the life of an MNA 

remedy. The illustration in Figure 3.2 provides an example 

where the off-site installation and activation of an irrigation 

well can result in movement of the ground-water plume in 

Figure 3.2 Potential effects of changes in ground-water flow direction on temporal trends in contaminant concentrations (USEPA, 2003).
47

response to this new ground-water withdrawal from the 

aquifer. In this situation, a decrease in contaminant concentration at a well screen originally placed within a portion 

of the plume with highest concentrations could be mistaken 

for evidence of an attenuation process. This misinterpretation in contaminant concentration trend can be prevented 

through periodic re-examination of the hydraulic gradient 

distribution within the monitored portion of the aquifer. In 

addition, these observations also provide guidance on how 

to adjust the location of monitoring points within the aquifer 

to insure that the entire plume is being monitored.

Elevations of surface-water bodies and ground-water elevation data periodically obtained from a network of wells 

and piezometers screened at appropriate depths within the 

contaminated aquifer units and surrounding units are essential elements of hydrologic analyses. Additional data for 

assessing ground-water flow include land use information 

such as the locations, rates, and schedules of irrigation; 

local precipitation data; and pumping rates and schedules 

for nearby wells. These data and the resulting estimates of 

hydraulic gradients are used in conjunction with interpretations of the hydraulic properties of subsurface materials 

and locations of contaminant sources to evaluate potential 

changes in contaminant loading, transport and attenuation 

rates, and transport directions with time. The evaluations 

should be three-dimensional in nature, including horizontal 

and vertical components of hydraulic gradients that result 

in three-dimensional contaminant transport. 

In many cases, frequent (e.g., weekly or monthly) monitoring of ground-water and surface-water elevations may be 

warranted, particularly during early phases of monitoring, to 

improve the characterization of ground-water flow patterns. 

In some cases, monitoring of these parameters on a very 

frequent basis using automated recording equipment may 

be needed to determine the effects of variability in recharge 

and discharge rates or locations. Once the effects are determined, the information would be used in the specification 

of appropriate long-term monitoring frequencies. These 

data may also indicate changes in hydraulic gradients that 

warrant more frequent monitoring of chemical parameters. 

Based on the results of such assessments, monitoring 

frequencies may be adjusted to adequately capture the 

fundamental features of the observed trends. In general, 

several years of monitoring data are often necessary for 

estimation of the site variability in the ground-water flow 

field. 

IIIA.4 Ground-Water/Surface-Water Interactions

Ground-water/surface-water interactions are of particular 

importance at sites where surface-water bodies are present. 

The hydrologic, as well as geochemical conditions, in 

areas where ground water discharges to or is recharged by 

surface water often differ markedly from those in the main 

body of the plume and may require intensive monitoring 

to determine the effect on remedial goals (Winter, 2000). 

Locations where ground water discharges to surface water 

may vary both temporally and spatially due to changing 

hydrologic conditions (Winter et al., 1998; USEPA, 2005a). 

Tools to characterize the hydraulic relationships between 

ground water and surface water include piezometers, 

devices for the direct measurement of ground-water flux 

and velocity, geophysical methods, and certain geochemical 

techniques. These tools may be used to define areas of 

plume discharge into the surface-water body, quantify rates 

of discharge, and to aid in the selection of monitoring point 

locations for determining the impact of the discharging 

water on the sediments and surface-water quality. Multilevel 

monitoring is generally required to characterize the 

interactions of ground water and surface water features. 

For situations where the contaminant plume intersects the 

transition zone between ground water and surface water, 

information on the distribution of ground-water discharge/

recharge can be used in the selection of monitoring points to 

characterize chemical characteristics of the transition zone. 

Characterization of the transition zone can be employed to 

assess and demonstrate that contaminant concentrations in 

discharging ground water or accumulated in surface water 

sediments do not negatively impact human or ecological 

receptors. Additional discussions of methods for hydrologic 

characterization of ground-water/surface-water interactions 

and monitoring of the potential for contaminated groundwater discharge to surface water may be found in various 

sources, including USEPA (2000a).

IIIA.5 Hydrogeologic Data Interpretation

The data obtained during hydrogeologic characterization 

are used in conjunction with information on contaminant 

sources, distribution, and behavior; redox conditions; and 

possible receptors to develop a conceptual model describing site conditions (Figure 3.3). A three-dimensional 

conceptual site model incorporating temporal changes is 

generally necessary to provide a framework for interpreting 

the site data, judging the significance of changes in site 

conditions, and predicting the range of future behavior of 

the source and plume. Understanding plume formation 

and behavior is the basis for evaluating whether an MNA 

remedy may be able to achieve site remedial goals within 

given time frames. Conceptual site models are expressed 

tangibly in text, maps (e.g., chemical isoconcentration maps 

and potentiometric surface maps), cross sections (e.g., 

hydrogeologic and chemical distributions), plots of temporal 

trends, and other graphical formats, and should be formulated in terms of mathematical calculations describing the 

plume and site. The conceptual model is a dynamic tool 

that is continually challenged, evaluated, and refined as 

new data are obtained. The data and analyses necessary 

for formulation of an adequate three-dimensional conceptual model for describing the effects of natural attenuation 

processes depend on site-specific conditions.

The development of quantitative models (i.e., mathematical 

models) based on the conceptual site model is often an 

important part of site characterization and remedy selection 

for MNA. These quantitative models may be as simple as 

equations for estimation of ground-water flow rates (e.g., 

Darcy’s law), or as complex as numerical models of groundwater flow and contaminant fate and transport. Such 

calculations are used to help understand site processes, 

locate monitoring points, estimate attenuation rates, and 

evaluate possible effects of different conditions on plume 
48

behavior. Quantitative models require particular types of 

data, and the data collection effort should be designed 

considering the requirements of the chosen model(s) and 

site-specific characterization objectives.

The conceptual site model for natural attenuation is the 

site-specific qualitative and quantitative description of the 

transport and fate of contaminants with respect to receptors 

and the geologic, hydrologic, biologic, and geochemical 

factors that control contaminant distribution (USEPA, 2003). 

The model expresses the understanding of the structure, 

processes, and factors that affect plume development and 

behavior. It is built on assumptions and hypotheses that 

have been tested using site-specific data.

IIIA.5.1 Attenuation Rate Estimates

The spatial and temporal distribution of the contaminant 

within the ground-water plume will depend on source location, the spatial distribution and velocity of water flow (or 

diffusion where advective flux is slow), and the abundance 

and biotic/abiotic reactivity of aqueous and solid phase 

biogeochemical components along the paths of water flow. 

The extent to which biogeochemical processes will cause 

contaminant attenuation during transport will depend on the 

relative rates of both fluid transport and chemical reactions 

(Morgan and Stone, 1985). For example, contamination 

will likely be negligible for systems in which the timescale 

for water flow within a hypothetical reaction volume is much 

shorter than the timeframe for significant reaction to take 

place. Conversely, significant attenuation may observed 

for contaminants in systems where reactions occur rapidly relative to the timescale for fluid transport. The latter 

situation is commonly assumed to apply to ground-water 

systems, but this needs to be confirmed during site characterization. This is particularly important for near-surface 

systems that may experience large variability in water flux 

and fluid velocities (e.g., Conant Jr., 2004).

Attenuation rate constant calculations can be an important 

tool for evaluating the feasibility of natural attenuation at 

a contaminated site, e.g., as part of the Tier II screening analysis. Specific applications identified in U.S. EPA 

guidelines (USEPA, 1999) include use in characterization 

of plume trends, as well as estimation of the time required 

for achieving remediation goals. As illustrated by Newell 

et al. (2002) different types of attenuation rate data may be 

obtained for a given site and careful consideration should be 

given to the appropriate use of the first-order rate constants 

derived from these data. 

The illustrations in Figure 3.4 provide two examples of 

attenuation rate constant data that may be collected within 

the plume. The overall extent of plume attenuation can 

be estimated by examining contaminant concentrations 

at a series of wells within the ground-water flow system 

(Panel A). In this instance, the influence of dispersion on 

contaminant concentrations as a function of distance must 

be determined through comparison with a conservative 

dissolved tracer. The duration of a plume at a given 

location can be estimated through analysis of contaminant 

concentration trends with time (Panel B). This can be 

used to estimate the time required to reach ground-water 

remediation goals. In both cases, the overall determination 

of whether the ground-water plume is expanding, showing 

relatively little change, or shrinking (in three dimensions) 

will depend on the analysis of multiple well transects 

(concentration vs. distance) or well points (concentration 

vs. time) in order to provide appropriate spatial coverage. 

Since the directions and rates of ground-water flow can 

vary in time, the determination of concentration-based 

rate constants may be subject to errors of a magnitude 

determined to be unacceptable for site characterization. 

In order to reduce this level of potential error, one may 

determine mass-based rate constants provided details on 

the spatial and temporal variability of hydraulic flux is also 

well characterized. Newell et al. (2002) should be consulted 

for further details on the application and limitations of the 

various approaches for estimation of rate constants.

IIIA.5.2 Contaminant Flux

While contaminant concentration is a determining factor 

for human or ecological risk, this metric does not provide 

a measure of contaminant mass and distribution within the 

system of interest. Determination of a contaminant mass 

balance is critical for determining changes in contaminant 

mass or speciation during subsurface transport. For organic 

contaminants, a mass balance calculation aids determination of whether contaminant degradation or sorption is 

occurring. For inorganic contaminants, a mass balance 

is required to assess changes in chemical speciation and 

mobility. The mass balance is inclusive of liquid and solid 

matrices (e.g., aquifer sediments) within the boundaries 

of the conceptual model. The distribution of contaminant 

mass is important with respect to defining remedial objectives and assessing remedial performance.

Contaminant flux (M) is defined as the product of contaminant concentration (C) in the mobile phase (water 

and mobile colloids) and the volumetric flow of the mobile 

phase (Q):

M = C * Q

Contaminant flux can be calculated for point locations or 

cross-sectional areas perpendicular to water flux depending 

on the level of heterogeneity in water flow or contaminant 

concentration distribution (Einarson and Mackay, 2001; 

Buscheck et al., 2004). For inorganic contaminants, this 

general equation represents the flux of all contaminant 

species at a given point in time. Since a mobile inorganic 

contaminant can change chemical form, it may be useful or 

necessary to further define contaminant flux for individual 

contaminant species relevant to site-specific conditions.

Mi

 = Ci * Q

Contaminant transport occurs along water flow paths in 

the subsurface. Therefore, the first step to determining 

contaminant flux is developing an understanding of system 

hydrology. With the establishment of a water budget for the 

site (i.e., water flux distribution), then a mass budget can 

be developed to establish contaminant flux across relevant 

system boundaries. 
49

Figure 3.3 Elements of a conceptual site model for monitored natural attenuation of inorganic contaminants (modified from USEPA, 2003).
50

IIIA.5.3 Source Term Characteristics

In addition to the geologic factors that influence groundwater transport, the chemical and physical characteristics 

of the source term can play a significant role in determining the physical dimensions of the ground water plume. 

The dimensions of the mobile plume may be impacted by 

characteristics of the contaminant source such as the total 

mass and rate of release of contaminant into the saturated 

zone within the aquifer. Active treatment of the source 

zone may also impact plume characteristics, particularly 

in cases where engineered remedies within the source 

zone introduce chemical reactants, stimulate microbial 

processes, or influence ground-water flow. Finally, the 

chemical characteristics of the contaminant source materials may impact the contaminant transport either directly 

through chemical reactions that influence aqueous speciation (e.g., Pb complexation by EDTA to form solution species) or indirectly by influencing ground-water chemistry 

(e.g., via biodegradation of organic co-contaminants such 

as fuel hydrocarbons or chlorinated solvents). In some 

instances, microbial reactions that are stimulated by the 

presence of organic contaminants may result in conditions 

beneficial to inorganic contaminant attenuation within the 

plume, e.g., generation of sulfate-reducing conditions that 

result in precipitation of metal sulfides. While this process 

may be beneficial to inorganic contaminant immobilization, the site investigation needs to include consideration 

of the long-term stability of the immobilized contaminant 

for a site at which the aquifer is generally oxic outside of 

the plume boundaries. This discussion serves to highlight 

the importance of factoring in evaluation of contaminant 

source characteristics and the management of the source 

area relative to down gradient plume behavior.

IIIB. Contaminant Quantification, Distribution 

and Speciation

Determination of the attenuation mechanism, assessment 

of aquifer capacity to attenuate contaminant mass, and the 

evaluation of the long-term stability of immobilized contaminants will necessitate analytical measurements conducted 

on aqueous and solid samples collected from within the 

plume. The types of measurements will include determination of total element concentrations as well as determination 

of aqueous and solid phase speciation of the contaminant 

and reactants involved in the attenuation reaction. The 

measurements will likely be applied to samples collected 

over a range of locations and times in order to adequately 

characterize the spatial and temporal variability of aquifer 

conditions within the plume. The following discussion will 

provide an overview of the types of measurements that 

may be required to evaluate the adequacy of site-specific 

attenuation reactions. The discussion will first address 

aqueous phase measurements followed by solid phase 

measurement techniques.

IIIB.1 Aqueous Characterization Approaches

Characterization of ground water will include assessment 

of the overall chemical conditions as well as the chemical 

speciation of the contaminant. Determination of the overall 

chemical conditions within the plume (e.g., redox status 

discussed in Section IIIC) sets the context for evaluating 

reactions that lead to contaminant attenuation. Chemical 

measurements to assess redox status of ground water can 

be achieved through a combination of field- and laboratorybased methods. Evaluation of the types and distribution 

of ions in ground water can adequately be addressed via 

laboratory measurements on properly preserved samples. 

Figure 3.4 Illustration of two approaches for determining attenuation rate constants within a contaminant plume. 

Critical assumptions underlie the applicability of these two analysis approaches: 1) analysis depicted in

Panel A presumes that a plume centerline exists and that five monitoring wells are sufficient to define it,

and 2) analysis depicted in Panel B presumes that the plume is stable. Both assumptions are to be verified through acquisition of supporting spatial and temporal data in order to reduce the uncertainty in rate

estimates.
51

These data are typically required to assess what type(s) 

of sorption reactions control contaminant immobilization. 

Likewise, determination of contaminant degradation 

products is needed to verify attenuation is occurring. An 

underlying assumption for all analysis types is that the 

approaches to sample collection and preservation are 

adequate to maintain the intact chemical characteristics 

of the ground-water sample. Insuring the adequacy of 

sample collection protocols can be realized, in part, through 

adherence to recommendations for well installation and 

low-flow well sampling procedures (USEPA, 1994; USEPA, 

1996). It is also recommended that the design of the well 

monitoring network consider temporal as well as spatial 

variability that may influence chemical measurements. 

For example, variations in ground-water levels may result 

in variations in ground-water redox chemistry. These 

considerations are explored more holistically in USEPA 

(2003), which should be consulted both from the perspective 

of designing the initial ground-water characterization plan 

as well as the long-term performance monitoring plan. 

In general, a variety of water samples should be collected 

at the same time to allow analysis for metals, common 

anions, alkalinity, total dissolved carbon, dissolved oxygen, 

and pH. In some systems, assessment of oxidation-reduction potential (ORP) using a Pt electrode may be desirable 

for qualitative assessment of variability between screened 

intervals within the plume. For metals and anions, the use 

of ultra-clean sample bottles is essential to avoid artifacts, 

and these can be checked through the inclusion of trip 

blank samples consisting of deionized water or tap water 

in the same types of containers. 

For preservation, samples for metal analysis are usually 

acidified immediately upon collection to pH 2 using either 

HNO3

 or HCl to avoid precipitation of oxide phases. Samples for anion analysis are not treated with acid, as this may 

swamp the anion of interest during the ion chromatography 

analysis. Subsamples of the same sample taken for anion 

analysis can be used for alkalinity determinations, as this 

quantity is robust. Measurements of dissolved oxygen 

and pH are best taken in the field using a closed flow cell 

instrumented with electrodes for in-line measurement. For 

most samples (especially total dissolved carbon), chilling to 

4 °C as soon as possible after collection is useful, although 

changes of temperature for dissolved gas measurements 

are not recommended. Methods used for determination 

of total concentrations of metals in aqueous samples are 

usually the same as used for destructive analysis of solids with elimination of the digestion step. Oxidation-state 

determinations can be made on the basis of reactions 

with chromophores [e.g., s-diphenyl carbazide for Cr(VI), 

phenanthroline or ferrozine for Fe(II)] and determination of 

absorbance in the UV-Vis region of the spectrum. Field kits 

and ion-selective electrodes are available for many of these 

determinations as well as for pH and dissolved oxygen. 

Common anions are usually determined in the laboratory 

using ion chromatography or capillary electrophoresis.

Colorimetry involves specific complexation of the atom 

of interest in solution by a strongly absorbing compound 

having a high absorptivity at an accessible wavelength 

(i.e., a chromophore) and subsequent measurement of the 

absorptivity of the solution. Atomic absorption, emission, 

and fluorescence rely on the same electronic transitions 

involving valence-shell electrons of unbound atoms. With 

absorption, the amount of light lost in exciting electrons 

into unbound states is measured, whereas with emission 

and fluorescence, the amount of light emitted when these 

electrons return to the ground state is measured. Emission 

and fluorescence differ in the manner by which the electrons are excited—emission involves thermal excitation of 

the sample by injection into a flame (flame photometry) or 

a plasma (ICP-AES), whereas fluorescence uses photon 

excitation. The sensitivity of colorimetry, AAS, AES, and 

AFS are roughly comparable (ppm to ppb), although ICPAES has the advantage of being able to analyze multiple 

elements more easily with fewer matrix interferences and 

across a wider concentration range than the other techniques. With ICP-MS, the masses of individual atoms are 

measured using a mass spectrometer, the intake of which 

is coupled to a plasma where the atoms are thermally 

ionized. This technique has many advantages including a 

wide dynamic range, ability to analyze all elements heavier 

than He, few matrix or spectral interferences, and the ability to measure isotopic ratios. Specific recommendations 

on contaminant speciation measurements and supporting 

ground-water chemistry data are provided within individual 

non-radionuclide contaminant chapters included in Volume 

2 of this document.

IIIB.1.1 Filtration

While regulatory requirements stipulate that unfiltered 

ground-water samples be analyzed to support regulatory 

decisions at a contaminated site, it may be necessary to 

also collect filtered samples to help interpret the process(es) 

controlling contaminant mobility. The use of 0.45 μm pore 

size filter paper is common as an arbitrary cutoff point to 

differentiate between dissolved and particulate phases in 

water samples. However, caution is recommended when 

using this approach, particularly for Fe and Al and other 

elements that may be associated with Fe or Al particles 

(including associated contaminants) that could pass through 

0.45 μm filter membranes. The use of filter membranes 

with pore sizes of 0.1 μm or less will generally provide a 

better assessment of the dissolved vs. particulate load of 

ground water.

Analytical methods have typically used 0.45 μm filters to 

differentiate between dissolved and particulate phases. If 

the intent of such determinations is an evaluation of truly 

dissolved concentrations, which would be important for 

geochemical modeling purposes, the inclusion of colloidal 

material less than 0.45 μm will result in incorrect values. 

Conversely, if the purpose of sampling is to estimate ‘mobile’ 

contaminant species in solution, significant underestimations of mobility may result, due to colloidal facilitated transport by particles, which are filtered out by 0.45 μm filtration. 

Kim et al. (1984) found the majority of the concentrations of 

rare earth elements to be associated with colloidal species 

that had passed a 0.45 μm filter. Wagemann and Brunskill 
52

(1975) found more than a two-fold difference in total iron 

and aluminum values between 0.05 and 0.45 μm filters of 

the same type. Some aluminum compounds were observed 

to pass through a 0.45 μm filter, but were retained on a 

0.10 μm filter (Hem and Roberson, 1967). DeVitre et al. 

(1987) found approximately 35% of the particulate bound 

manganese in the 0.015 to 0.10 μm size fraction in anoxic 

lake waters. Kennedy and Zellweger (1974) found errors 

of an order of magnitude or more in the determination of 

dissolved concentrations of aluminum, iron, manganese 

and titanium using 0.45 μm filtration as an operational 

definition for “dissolved”. Sources of error were attributed 

to filter passage of fine-grained clay particles. De Mora 

and Harrison (1983) provided an excellent review on the 

subject of physical separation techniques for trace metal 

speciation.

During sample collection in anoxic or suboxic systems, iron 

oxidation and precipitation may occur prior to filtration and 

result in the removal from solution of previously dissolved 

species due to instantaneous sorption by the precipitate 

(Puls and Eychaner, 1990). Filter loading and clogging of 

pores with fine particles may also occur, introducing filtration 

errors due to reductions in effective pore size (Danielsson, 

1982). Sheldon and Sutcliffe (1969) found that virtually all 

filters remove particles smaller than the stated pore size. 

In experiments with seawater and latex particles, using 

light scattering techniques, Johnson and Wangersky (1985) 

demonstrated that a high proportion of materials dispersed 

at sizes smaller than the filter pore size will interact with the 

filter surface. These interactions are dependent upon size, 

particle concentration, colloid surface chemistry, electrolyte 

concentration and composition, nature and concentration 

of adsorbents, chemical properties of the filter surface, and 

the frequency of collisions of dispersed particles with the 

filter surfaces. 

Contaminants may exist as dissolved species, precipitated 

solids, polymeric species or be adsorbed to inorganic or 

organic particles of colloidal dimensions. Based on the 

above discussions, the filtration of ground water samples 

for metal analyses using 0.45 μm filters may not provide 

accurate data either for geochemical modeling or for contaminant migration estimates. Some mobile species are 

likely to be removed by 0.45 μm filtration prior to chemical 

analysis, while other particulate-associated metals will pass 

through the filters and be incorrectly considered ‘dissolved’. 

A principle objective in sampling to test a geochemical 

speciation model is to obtain estimates of the free ion activities of the major, minor, and trace elements of interest. 

Since there are relatively few easily performed analytical 

procedures for making these estimates, an alternative is to 

test the analytically determined dissolved concentrations 

with model predictions, including both free and complexed 

species.

Collection of ground-water samples and measurement of 

their chemical properties will be necessary irregardless 

of whether degradation or attenuation is the primary attenuation mechanism. Characterization of aquifer solids 

may be necessary to support evaluation of degradation 

processes, but primarily when knowledge of specific microbial processes or identification of abiotic solid phase 

reductants is needed to constrain attenuation rate and 

capacity evaluations. In contrast, for contaminants in which 

immobilization is the primary (or only) viable attenuation 

mechanism, collection and characterization of aquifer solids 

is a specific requirement.

IIIB.2 Solid Phase Characterization 

Approaches

Determination of aquifer solids mineralogy and solid phase 

contaminant speciation is critical to identification of the contaminant immobilization process. Contaminant speciation is 

inclusive of oxidation state of the immobilized contaminant 

as well as the solid phase component(s) with which it is 

associated. Determination of contaminant solid phase 

speciation can be approached in two ways, structurally and 

operationally. Structural determination uses instrumental 

procedures (spectroscopy, microscopy) to identify and 

quantify the discrete phases present. This information, together with accumulated knowledge of their thermodynamic 

properties and the rates at which they precipitate/adsorb 

and dissolve/desorb, is employed to assess the capacity of 

the aquifer to sustain contaminant attenuation and evaluate 

the long-term stability of immobilized contaminants. The 

challenges associated with the structural approach include 

1) difficulty in identifying specific phases at concentrations 

near or below instrumental detection limits, 2) need for good 

sampling statistics (e.g., a large number of micron-sized 

measurements are needed to ensure that the site is adequately represented), 3) inadequacy of available thermodynamic and kinetic data to evaluate reaction mechanisms, 

and 4) cost of instrumentation and model applications that 

may be employed during evaluation of aqueous and solid 

phase chemical data.

The operational approach, on the other hand, classifies 

contaminant form solely in terms of reactivity. Contaminated materials are contacted with solutions that simulate, 

in a short period of time, subsurface conditions expected 

over much longer time intervals. The amount of contaminant released by the material as a result of this contact is 

measured. In some instances, the rate of release is also 

quantified either by in-situ measurements or successive 

extractions. Examples of the operational approach include 

chemical extractions using a variety of solutions and procedures, and bioavailability studies that measure uptake 

by organisms directly. Sequential chemical extractions, in 

which contaminated materials are treated with successively 

harsher solutions, have found wide use in site characterization and several standard methods are available for specific 

situations (Tessier et al., 1979; Yanase et al., 1991). While 

highly relevant, bioavailability studies are rarely conducted 

because of the high cost and long time periods required. 

Operational determination of contaminant form avoids many 

of the problems associated with the structural approach, but 

still requires the development of substantial and expensive 

databases that correlate the amounts and rates of contaminant release during laboratory contact studies with actual 

contaminant reactivity under field conditions.
53

IIIB.2.1 Sampling and Fractionation

Collection of samples is the first step in determination of 

contaminant distribution within the source area and down 

gradient plume. The most important consideration is that 

the samples collected give a fair representation of the 

size and extent of the contamination. Enough samples 

must be collected to provide statistical certainty to any set 

of results. As soils and sediments are heterogeneous, 

minimum sample sizes of about 50 g are needed at each 

sampling point. These are mixed thoroughly before taking 

1- to 10-g subsamples for analysis. In some instances, 

mixing of samples from different sampling points to provide 

composite samples can be done to decrease analytical 

costs, but this practice also tends to decrease the precision 

with which the geographical extent of the contamination 

can be determined. To avoid the possibility of bias, and 

to retain the ability to interpolate results, sampling points 

should be laid out on a three-dimensional grid that takes 

into account any pre-existing knowledge of the underlying 

stratigraphy. Samples may be collected in several rounds, 

starting with a coarse grid to identify the overall extent of 

the contamination and possible “hot” spots. Successive 

sampling rounds using finer grid sizes on subsets of the 

original coarse grid can be used to provide further data 

about regions of interest.

Aquifer samples can be collected using a variety of methods. 

In all cases, however, the goal will be to collect materials 

to allow for lithologic logging, to provide representative 

samples for laboratory investigations and for submission 

to analytical laboratories. The samples must meet the 

appropriate data quality objectives as identified in the project-specific Quality Assurance Plan (QAP). In all cases, 

aquifer materials should be collected with local, State, and 

Federal requirements in mind.

The procedures to follow for collection of soil and sediment 

samples will depend on the degree to which environmental 

availability is to be assessed. Aquifer materials can be collected, for example, using hollow-stem auger or hydraulic 

percussion methods. These are generally well-tested 

methods applicable to a wide variety of environments. 

With these methods various types of sample liners may be 

used, such as plastic or brass sleeves. The sleeves can 

be cut and capped to preserve materials and facilitate their 

transport. If oxidation-reduction processes are believed to 

be an important component of attenuation mechanism(s), 

special attention must be given to preserving the redox 

status of materials after they are retrieved from the subsurface. For example, if anoxic materials are collected they 

must be frozen after collection or stored in evacuated or 

inert-gas-purged containers in order to preserve primary 

mineralogy (USEPA, 2002; USEPA, 2006).

For total analyses, little care need be taken and the samples, 

once collected, can be air-dried and stored before analysis. 

The availability of many metals, however, depends greatly 

on their oxidation state, and this may change with time in 

storage due to microbial activity or exposure to air. Drying 

of samples can also affect the availability of some elements, 

e.g., due to irreversible (or slowly reversible) changes in 

the structure of illitic and vermiculitic clay minerals during 

drying. A realistic assessment of current availability, therefore, will attempt to make the measurement as soon after 

collection as possible using samples that have been stored 

under conditions that maintain their redox and moisture 

status at the levels seen in the field (USEPA, 2006). The 

best procedures for preserving samples thus involve, at 

a minimum, storage of field-moist samples at 4 °C in airtight plastic bags from which the air has been squeezed. 

In instances where oxidation state is critical, injection of 

N2

(g) or Ar(g) into the bore hole, and collection of cores 

in thick plastic cylinders that are immediately capped and 

transferred into anoxic chambers for further processing or 

storage is advised. Lowering the temperature to 4 °C slows 

microbial activity substantially. Freezing the samples may 

result in lysing of microbial cells and can create a different 

organic mixture in the soils than was present at the time 

of sampling.

In preparation for actual analysis, some particle-size fractionation may be needed for easy sample handling and 

enhanced sensitivity. As a result of surface-area considerations, the most reactive fraction of soils and sediments 

is that having particle sizes less than about 50 μm (i.e., 

silt and clay). Gravel (particles > 2 mm) and cobbles can 

generally be removed by sieving with little or no impact on 

the analytical results. The sand fraction (50-2000 μm) is 

generally left intact, as its removal by wet-sieving creates 

more problems than it solves. Careful records need to be 

kept of the mass and volume of the fractions removed so 

that analytical results on the remaining fractions can be 

scaled to field conditions. Fractionation on the basis of 

properties other than particle size can be done to concentrate specific minerals, and this may prove beneficial in 

some instances (Laird and Dowdy, 1994).

IIIB.2.2 Total Amount

Determination of the total quantity of a metal contaminant 

present is generally performed on “bulk” sample sizes of at 

most a few grams. In instances where the total amount of 

contamination is very low and detection limits are encountered, analysis by electron or X-ray microscopic techniques 

can be performed in which specific particles containing the 

contaminant are identified. With the microscopic approach, 

sample sizes of μg to mg are common, and scaling to field 

concentrations is difficult. 

The bulk analysis techniques are distinguished on the basis 

of whether they are sample destructive or nondestructive. 

The sample-destructive techniques require digestion of the 

sample using strong acid or base to destroy the original 

compounds present and release the elements in a soluble 

state. Once in the soluble state, the concentration of the 

contaminant is determined by spectroscopic means after 

nebulization or injection into a flame or plasma connected to 

a suitable detector. The most common detection techniques 

include colorimetry, atomic absorption spectrometry (AAS), 

atomic emission spectroscopy (AES) using either flame 

(photometry) or an inductively coupled plasma (ICP-AES), 

atomic fluorescence spectroscopy (AFS), and inductively 

coupled plasma mass spectrometry (ICP-MS). 
54

Colorimetry involves specific complexation of the atom 

of interest in solution by a strongly absorbing compound 

having a high absorptivity at an accessible wavelength 

(i.e., a chromophore) and subsequent measurement of the 

absorptivity of the solution. Atomic absorption, emission, 

and fluorescence rely on the same electronic transitions 

involving valence-shell electrons of unbound atoms. With 

absorption, the amount of light lost in exciting electrons 

into unbound states is measured, whereas with emission 

and fluorescence, the amount of light emitted when these 

electrons return to the ground state is measured. Emission and fluorescence differ in the manner by which the 

electrons are excited—emission involves thermal excitation 

of the sample by injection into a flame (flame photometry) 

or a plasma (ICP-AES), whereas fluorescence uses photon excitation. The sensitivity of colorimetry, AAS, AES, 

and AFS are roughly comparable, although ICP-AES has 

the advantage of being able to analyze multiple elements 

more easily with fewer matrix interferences and across a 

wider concentration range than the other techniques. With 

ICP-MS, the masses of individual atoms are measured 

using a mass spectrometer, the intake of which is coupled 

to a plasma where the atoms are thermally ionized. This 

technique has many advantages including a wide dynamic 

range (pptr to ppm), ability to analyze all elements heavier 

than He, few matrix or spectral interferences, and the ability 

to measure isotopic ratios.

The nondestructive analysis techniques generally rely on 

inner-shell electronic transitions or nuclear transitions for 

elemental identification and thus are less reliant on decomposition of the sample into individual atoms. Nevertheless, a 

homogeneous sample consisting of particle sizes of <2 μm 

and presenting a flat surface is needed for best results. 

Thus, although the sample is not completely destroyed in 

preparation, it is altered by grinding to achieve the particle 

size needed. The most common of these techniques is Xray fluorescence (XRF), which can use high-energy photons 

from an X-ray tube or radioactive source to excite electrons 

in the atoms of interest. The light emitted by these atoms 

upon return to the ground state is then detected using an 

X-ray detector such as a gas-filled proportional counter, 

scintillation detector, or solid-state semiconductor detector. X-ray emission spectroscopy is similar to XRF except 

that it uses high-energy particles (electrons, protons, or 

alpha particles) to excite the atoms rather than photons. 

In general, detection limits are in the ppm range for solids 

although X-ray emission spectroscopy is more sensitive 

for lighter elements (z<30) and X-ray fluorescence is more 

sensitive for heavier elements (z>45) (Amonette and Sanders, 1994).

Microscopic methods of elemental analysis are generally 

more expensive than the bulk techniques, but can prove of 

great use in identification of specific contaminant phases 

and their associations with minerals. Electron microscopy 

images particles through scattering of an electron beam. At 

the same time, the electron beam is stimulating the emission of X-rays by the sample and detection of these (as in 

X-ray fluorescence) allows simultaneous multielemental 

analysis of the material being imaged. Scanning electron 

microscopy (SEM) can be used with samples of any thickness, whereas transmission electron microscopy (TEM) 

requires preparation of samples on the order of 30-150 nm 

thick often in the form of thin sections. Conventional SEM 

and TEM require high vacuum for the analysis, due to the 

short pathlength of electrons in air, but new developments 

of the environmental SEM allows analysis of samples at 

pressures approaching ambient thus allowing observation of 

wet specimens for short periods of time. With the development of synchrotron X-ray sources, a microscopic technique 

based on X-ray fluorescence is available for high-value 

specimens. X-ray microscopy (XRM) has spatial resolution 

near 1 μm, and because of the very high photon fluxes available can achieve detection limits in the ppb range, better 

than any other X-ray technique. The XRM technique can 

be coupled with X-ray absorption spectrometry (XAS) to 

determine oxidation states and local structural information. 

Although primarily a research tool at the moment, XRM 

may prove quite useful for phase identification of critical 

samples when other techniques fail.

IIIB.2.3 Structurally Defined Form

One approach to the determination of environmental 

availability relies on identification of the discrete contaminant-bearing phases that are present and combining this 

information with phase-relevant thermodynamic or kinetic 

data using a geochemical model. Identification and quantification of the discrete phases present is gained through a 

combination of elemental, structural, and in some instances, 

solubility analysis. Techniques for structural analysis measure the arrangement of atoms both local to the contaminant 

and in extended structures. The combination of structural 

and elemental information uniquely determines a thermodynamic phase with specific properties. 

The primary technique for structural analysis is X-ray diffraction (XRD), which relies on measuring the coherent 

scattering of a collimated beam of X-rays by the sample as a 

function of angle to the beam. The technique requires some 

degree of repetitive structural unit having dimensions on the 

order of the wavelength of the X-rays. Scattering from parallel planes of atoms separated by regular spacings yields an 

interference or diffraction pattern. The spacings at which 

constructive interference occurs (d) yield high intensity in 

the diffraction pattern and can be calculated from the angle 

of incidence (θ) and the wavelength of the X-ray (λ) using 

the Bragg equation, d = nλ/2sinθ, where n is an integer. 

The set of d-spacings and relative intensities measured for 

a compound are unique and can be compared with those for 

thousands of other compounds stored in a large database. 

In soils and sediments, however, mixtures of compounds 

occur and overlapping patterns for different compounds 

can make identification difficult in many instances. Also, 

XRD is sensitive down to about 1-5% by weight, and thus 

is not well-suited for identification of trace phases typical of 

contaminants in soils and sediments. To some extent, these 

drawbacks can be overcome by fractionating the sample 

to eliminate interfering compounds and to concentrate the 

contaminant-bearing phases. Diffraction also occurs with 

electron beams, as these have wave properties (albeit with 

shorter wavelengths than X-rays). Electron diffraction is 

used in TEM analysis to identify phases at the same time 
55

they are being imaged and their elemental compositions 

determined by X-ray emission. Diffraction techniques do 

not work well for poorly ordered phases, as these phases 

yield very broad peaks that have little value for identification purposes. 

Vibrational spectroscopic techniques can be used in some 

instances to identify poorly ordered phases, as these 

methods detect molecular vibrations that depend on bond 

energies and atomic masses rather than long-range atomic 

order. The two major types of vibrational spectroscopy 

are infrared absorption and Raman scattering. Infrared 

absorption spectroscopy requires a long-wavelength (ca. 

1-100 μm) source of light and measures the fraction of 

this light that is absorbed by the sample. Water absorbs 

strongly in much of this region, and so samples are best 

analyzed when dry, unless internal reflectance techniques 

are used. The long wavelengths used, however, are well 

suited for interferometry in which the incident beam is split 

in two and recombined after delaying one beam by a known 

path length. The resulting interference spectrum, collected 

as a function of path length difference, is then converting 

using a Fourier transform to yield an absorption spectrum 

in terms of wavelength. Because the entire spectrum is 

collected at all times, rather than scanning through each 

wavelength individually, the Fourier-transform approach 

(i.e., FTIR) is the most efficient means of collecting an 

infrared absorption spectrum.

Raman scattering spectroscopy measures the loss or gain 

in energy of monochromatic light as a result of interactions 

with molecular bonds. In theory, incident light of any wavelength can be used to measure these changes, and so some 

of the limitations of infrared absorption spectroscopy are 

avoided, such as the need to study dry samples. However, 

Raman scattering yields a weak signal, and fluorescence 

from the sample may interfere when incident light in the 

visible region is used. The recent development of Fouriertransform Raman spectroscopy using laser light in the near 

infrared region (ca 1 μm wavelength) has provided enough 

incident light intensity with minimal fluorescence to make 

Raman practical for many samples. 

Vibrational spectroscopic techniques, while sensitive to local bond strengths, are rarely as definitive as XRD because 

the same types of bonds are found in most inorganic compounds (e.g., oxide, hydroxyl, sulfate, carbonate, etc.) and 

the differences among compounds are often subtle. This 

approach, however, can be used to positively identify solid 

solutions of poorly ordered materials, such as (Fe,Cr)(OH)3

(Amonette and Rai,1990), based on shifts in bond energies 

due to substitution. Vibrational techniques also find use for 

identifying orientation and bonding of sorbed contaminants 

that exist in ground water as oxyanions (e.g., arsenic; Eggleston et al., 1998).

Under certain circumstances, solubility can be used to 

determine or at least confirm the environmentally relevant 

phase controlling the availability of an inorganic contaminant. These circumstances require collection of equilibrium 

solubility data across a range of pH (or some other primary 

solubility-controlling variable), in which the concentrations 

of the contaminant as well as other key species involved in 

formation of aqueous complexes with the contaminant are 

measured. Based on these data, and known thermodynamic solubilities of the putative contaminant phase, a solubility 

diagram can be constructed and compared with that for 

known phases. It is critical that equilibrium be attained or 

thermodynamic calculations will not be valid. Thus systems 

that approach equilibrium slowly may not be suitable. Rai 

et al. (1984) reviewed this approach and its applicability to 

twenty-one environmentally relevant elements. 

IIIB.2.4 Operationally Defined Form

A more direct approach to assessing environmental availability of a contaminant, involves determining the conditions 

under which it can be dissolved and therefore mobilized 

into the ground water. This empirical approach presupposes nothing about the chemical form of the contaminant 

and thus has little predictive value in the event that ground 

water conditions change. Nevertheless, it offers much in 

the way of addressing economically the degree to which 

a total concentration of a contaminant poses an environmental risk.

The primary technique used is that of sequential selective extractions, with each successive extracting solution 

offering a harsher solution environment than the one that 

preceded it. The general assumption is that the earlier in 

the sequence that the contaminant is released, the higher 

is its environmental availability. Sequential extraction approaches are discussed in more detail in Sections IIIB.2.4.1 

and IIIB.2.4.2.

The sequential (selective) extraction approach can be 

improved by incorporating measurements of the rate of 

release during each step in the procedure. This can be 

done by successive extractions for shorter time periods 

using the same reagent, by measurement of small aliquots 

taken from the extraction solutions at different times during 

the extractions, or by a continuous-flow extraction using a 

specially designed flow cell similar to that of Wollast and 

Chou (1985) used in mineral weathering studies. Despite 

the apparent need, the kinetic approach has not been 

widely applied to estimation of environmental availability 

of contaminants, although it has been suggested for use 

in assessing the availability of U in soils (Amonette et al., 

1994b).

IIIB.2.4.1 Sequential Extractions

Sequential extraction methods will perhaps be the most 

practicable approaches for demonstrating contaminant 

partitioning in the solid phase and for providing information for substantiating proposed attenuation mechanisms. 

These methods can be particularly useful because they 

provide quantitative information on the capacity of a given 

material to attenuate inorganic contaminants. They are 

also advantageous because large numbers of samples can 

be analyzed and compared, unlike methods that involve 

spectroscopy and microscopy. However, as a cautionary 

note, sequential extraction methods are not without limitations. It must be acknowledged that results of sequential 

extraction tests are operationally defined and users of these 

tests must be aware of potential artifacts, as will be pointed 

out in the following discussion.
56

Sequential extraction procedures consist of subjecting a 

given quantity of soil or aquifer sample to a series of increasingly aggressive reagents under specified conditions. An 

underlying assumption is that the release of components in 

earlier extractions implies a higher potential for environmental mobility than components released later in the extraction 

sequence. A conceptual model of the sequential extraction 

approach is that solid materials consist of specific mineral 

fractions that can be extracted selectively by using appropriate reagents and experimental conditions. As mineral 

fractions are selectively dissolved, any element that they 

contain will release into solution and in this way the “speciation” of particulate trace metals can be determined. If it is 

determined that the majority of a contaminant of concern is 

present, for example, in the operationally-defined carbonate fraction, then geochemical modeling efforts might be 

appropriately focused on the factors that govern long-term 

stability of carbonate minerals. 

Perhaps the most often followed or modified chemical 

extraction procedure is that of Tessier et al. (1979). Other 

approaches are outlined, for example, in Dragun (1988) and 

Yong et al. (1993). In the Tessier et al. (1979) method, five 

distinct extraction procedures are used to recover metals 

from the following sediment/soil fractions: 1) exchangeable sites; 2) carbonate minerals; 3) metal oxyhydroxides; 

4) organic matter; and, 5) the residual fraction (Table 3.1). 

Metals associated with loosely bound sites (weak electrostatic attraction) are released by extraction in concentrated 

salt solutions. Divalent metal chloride salts prepared at a 

concentration of 1 molar are often used in this first step 

(e.g., CaCl2 or MgCl2). After extraction for a set period of 

time and mixing rate, the supernatant solution is analyzed 

using spectroscopic methods and the concentration of an 

element in solution is related back to a mass fraction of that 

element associated with operationally defined exchangeable sites. Metals associated with carbonate minerals are 

removed in a second step by wet chemical extraction in a 

sodium acetate-acetic acid buffer solution. This buffer (pH 

5) is effective in dissolving calcite, an abundant carbonate 

mineral found in natural systems. Iron and manganese 

oxides are removed with hot hydroxylamine hydrochloride 

solution. Organic compounds, such as humic acids and 

fulvic acids, are targeted with acidic, oxidizing reagents 

(hydrogen peroxide plus nitric acid; note that other published 

extraction methodologies employ basic conditions to “selectively” remove organic matter, for example, 3 M KOH). The 

residual fraction is typically determined from a concentrated 

nitric acid digestion or total analysis (by complete digestion 

or by X-ray fluorescence spectroscopy) after subtracting out 

the fraction of metal extracted in steps 1 through 4. The 

residual fraction is considered to represent the fraction of 

metals present in tightly bound matrices, for example, in 

aluminosilicate matrices that are not effectively dissolved 

with any of the reagents used in steps 1-4.

This extraction scheme has been adapted for testing and 

validation against natural matrix standards by the Radioactivity Group of the National Institute of Standards (NIST) 

Ionizing Radiation Division (http://physics.nist.gov/Divisions/

Div846/Gp4/environ.html). This work represents an effort 

to support metrology improvements in the radiochemistry 

community through research and development of low-level 

radionuclide Standard Reference Materials (SRMs) and 

Table 3.1 Sequential extraction procedure of Tessier et al. (1979).

Step Target Fraction Chemicals/Conditions

Step 1: 

Exchangeable Ionically bound metals and metalloids

1 M magnesium chloride 

(pH = 7.0) 

1 hour, room temperature

Step 2: 

Carbonates

Metals/metalloids with carbonate minerals 

(e.g., calcite, dolomite)

1 M sodium acetate plus 

acetic acid (pH = 5.0) 

4 hours, room temperature

Step 3: 

Fe and Mn oxides/

hydroxides

Metals/metalloids with iron and 

manganese hydroxides

0.04 M hydroxylamine hydrochloride in 25% 

(v/v) acetic acid (pH ~ 2) 

6 hours, 96°C

Step 4: 

Organic matter

Metals/metalloids bound to natural 

organic matter (e.g., humic and fulvic 

acids)

0.02 M nitric acid, 18% hydrogen peroxide 

(pH = 2) 

2 hours, 85 °C

Step 5: 

Residual Aluminosilicate minerals, pyrite 16 N nitric acid 

2 hours, 140 °C

Notes: Initial sample weight is generally 1 gram. This initial mass is used to report all metals fractions in terms of metal weight per gram of dry

sediment. Samples are generally dried (aerobically or anaerobically) to a constant mass, gently ground and homogenized in an agate mortar. Samples should be constantly agitated during each extraction step. Extractions are conveniently carried out using polypropylene or Teflon

centrifuge tubes. Initial and final pH values should be measured and recorded. Supernatant solutions are collected and analyzed by atomic

absorption spectroscopy, inductively coupled plasma emission spectroscopy, or other suitable methods. Between steps, samples are centrifuged

(typically 5000 to 10000 rpm, 30 minutes) and rinsed with deionized water (5-10 mL). 
57

measurement quality assurance. The Radioactivity Group 

has undertaken a systematic study to test the relative selectivity of a Tessier-based extraction scheme for the speciation 

of a suite of radionuclides and stable elements (Schultz et 

al. 1998a; 1998b; 1998c; 1999). These tests have been 

conducted with a suite of natural matrix standards (soils 

and sediments) developed by NIST for the purpose of 

validating radiochemical methods. While the adoption of 

this extraction scheme may not be suitable for all sample 

matrices (see discussion below), the development of a 

uniform reference database relative to the performance 

for a given extraction scheme provides a valuable starting 

point for adapting an extraction scheme(s) better suited for 

site-specific conditions. The availability of certified SRMs 

for testing extraction schemes is critical for assessing data 

quality across analytical laboratories that may be utilized 

to support site characterization.

In sulfate–reducing systems, the iron sulfides mackinawite 

(FeS) and pyrite (FeS2) may be potential hosts to a variety 

of metals. Huerta-Diaz and Morse (1990) presented a sequential extraction method for the determination of metal 

partitioning in iron sulfides. Their method involves the 

sequential leaching of samples using 1 M HCl to recover 

a “reactive” fraction (FeS; equivalent to acid volatile fraction 

discussed below), 10 M HF to recover the silicate fraction, 

and finally concentrated HNO3

 to recover the pyrite fraction. 

Moore et al. (1988) investigated the partitioning of metals 

in reducing sulfidic systems using sequential extraction 

techniques. They chose to use a three-step procedure: 

1) 0.25 M hydroxylamine hydrochloride in 25% acetic acid 

to remove metals bound to oxyhydroxides of Fe and Mn 

and carbonates; 2) 0.1 M sodium pyrophosphate at pH 10 

to recover the organic fraction; and 3) the sulfide fraction 

was recovered using potassium chlorate plus hydrochloric acid. These fractions were compared with total metal 

concentrations determined by digestion of samples in HF 

plus perchloric acid.

Numerous reagents and extraction schemes have been 

developed for iron minerals in soils and sediments (e.g., 

Mehra and Jackson, 1960; Chao and Zhou, 1983; Walker, 

1983; Canfield, 1989; Ryan and Gschwend, 1991; Kostka 

and Luther, 1994). Because of the importance of iron 

minerals for sequestering metals and metalloids in the 

environment, it is worthwhile to summarize the most commonly used wet chemical techniques for iron, and the oxide 

and sulfide minerals dissolved using these various reagents 

(Table 3.2). Slight variations in technique can have profound influences on the outcome of chemical extractions, 

for example, whether 1 M or 4 M hydrochloric acid is used 

(e.g., Chao and Zhou, 1983). Consequently, in all cases 

where sequential extraction methods are employed it will 

be necessary to clearly document the types of reagents 

used, extraction times, temperature, and laboratory procedures. Validation of laboratory procedures by spiking 

Table 3.2 Summary of reagents used to selectively dissolve iron oxides and sulfides.

Extractant Composition pH Time (h) Fe minerals Extracted

Ascorbate1 0.17 M sodium citrate; 0.6 M sodium 

bicarbonate; 0.023 M ascorbic acid 8 24 ferrihydrite

HCl1-3 0.5–6 M hydrochloric acid <2 1 ferrihydrite mackinawite

Hydroxylamine 

hydrochloride1- 4

0.04 M hydroxylamine hydrochloride in 

25% (v/v) acetic acid, 96 °C <2 6 ferrihydrite mackinawite

Oxalate5 0.2 M ammonium oxalate/ 0.2 M oxalic 

acid 2.5 48 ferrihydrite lepidocrocite magnetite (partial) mackinawite

Dithionite1,2,6,7

0.11 M sodium bicarbonate/0.27 M 

sodium citrate, add 0.5 grams sodium 

dithionite at 80 °C, repeat dithionite 

addition

5-7 4

ferrihydrite 

goethite (partial) 

hematite (partial) 

magnetite (partial)

Ti-citrate8 0.05 M Ti(III) citrate EDTA-bicarbonate 7 2

ferrihydrite 

goethite 

hematite (partial)

Chromous 

chloride9,10

1 M Cr(II) chloride in 0.5 M 

hydrochloric acid; reagent prepared 

using a Jones reductor

<2 1

pyrite 

mackinawite 

elemental sulfur

Notes: see references for experimental details; 1Kostka and Luther (1994); 2Canfield (1989); 3Chao and Zhou (1983); 4Tessier et al. (1979); 5Schwertmann (1964); 6

Mehra and Jackson (1960); 7Walker (1983); Ryan and Gschwend (1991); 9Zhabina and Volkov (1978); 10Tack et al. (1997).
58

extraction experiments with known quantities of minerals 

and assessment of recovery will be needed in most cases 

to evaluate and verify assumptions related to sequential 

extraction approaches (see, for example, Chao and Zhou, 

1983; Kostka and Luther, 1994; Keon et al., 2001). In this 

way, the results of sequential extraction experiments can 

depart from purely operational definitions and approach 

the point where such results can be confidently applied to 

assess mineral-contaminant associations.

Over the past 10 years, the European environmental research community has developed a three-step sequential 

extraction scheme, the so-called BCR method (formerly 

the Community Bureau of Reference, now referred to as 

The Standards, Measurements and Testing Programme of 

the European Commission; see Rauret et al., 1999). The 

extraction scheme has been evaluated using standard reference materials (CRB 601, sediment; CRM 483, sewage 

sludge amended soil) by multiple laboratories (Ure et al., 

1993; Quevauviller et al., 1997; Rauret et al., 1999). In step 

1 of the BCR scheme, metals present in ionic forms, bound 

to carbonates, and in exchangeable forms are separated 

(Table 3.3). In step 2, metals bound to amorphous iron 

and manganese oxides are leached, while in step 3 metals 

bound to organic matter and sulfides are selectively dissolved (Table 3.3). The use of prescribed procedures and 

standard reference materials to verify performance is an 

important development and a potential advantage of the 

BCR extraction approach.

Assessments of environmental risk may be more suitably 

focused on the water soluble and exchangeable soil fractions, i.e., the most labile metal/metalloid forms released 

in the first step of the sequential extraction schemes described above. While total metals contents and pseudo 

metals contents (sum of all extractable forms) are valuable 

in defining the extent of metal buildup in contaminated 

soils and sediments, these fractions may be less useful for 

assessing environmental and ecological impacts (Gupta 

et al., 1996). Use of single-step extraction tests with salt 

solutions of CaCl2

 and NaNO3 for the determination of the 

most readily bioavailable metal fraction is discussed, for 

example, by Maiz et al. (1997) and Gupta et al. (1996). Although single-step procedures may be useful in developing 

models of ecological risk, these procedures will generally 

not be adequate in developing the necessary understanding of attenuation mechanisms and long-term contaminant 

behavior needed to adopt monitored natural attenuation as 

a site remedy.

In most cases, assessment and monitoring of natural attenuation processes for inorganic contaminants will require 

some type of sequential extraction procedure in Tier II. Tier 

II and Tier III sequential extraction efforts might be carried 

out to: i) further refine the conceptual model of natural attenuation; ii) increase the spatial resolution of metal partitioning data at a site; and, iii) validate extraction procedures 

by spiking extraction experiments with known quantities 

of reference minerals to assess method performance and 

test assumptions regarding solid phase partitioning. The 

selection of a sequential extraction procedure must take 

into consideration site-specific factors, such as physical soil/

aquifer characteristics, redox conditions, and contaminant 

type. Examples of sequential extraction approaches on a 

contaminant-specific basis are presented in the elementspecific chapters in Volume 2 of this document. It is possible 

that several procedures might be tried before an optimal 

method is selected at any site. Samples for sequential 

extraction tests should be collected near the source region 

and at points moving down gradient through the contaminant plume and at points past the down gradient and lateral 

plume fronts. Vertical resolution should cover the area most 

impacted by the contaminant plume but vertical sampling 

should also encompass subjacent and superjacent regions 

of the aquifer. These spatial data will be needed in order 

to develop a model of contaminant uptake along the flowpath. Sequential extraction procedures should be carried 

out prior to any active remediation and they may be a useful 

component of annual site monitoring activities.

IIIB.2.4.2 Sequential Extraction Considerations

Many examples of selective extraction recipes can be found 

in the literature. The choice of procedure will necessarily 

involve consideration of project objectives and site-specific 

details. Moreover, the pros and cons of using sequential 

extraction methods for determining metals speciation in the 

solid phase have been extensively discussed and debated 

in the literature (e.g., Kheboian and Bauer, 1987; Belzile et 

al., 1989; Nirel et al., 1990; Tessier and Campbell, 1991). 

Sequential extractions clearly provide a very practical 

methodology for getting at critical information about where 

metals reside in a sample that can be linked to bioavailability and geochemistry of contaminant metals. Further, 

a considerable amount of work has gone into verifying the 

selective extraction methods for specific elements of interest 

Table 3.3 BCR extraction scheme applied to 1 gram of 

sample.

Step Reagents/Conditions Target Fraction

Step 1

40 mL of 0.11 M acetic 

acid; 16 h at room 

temperature 

Ionic, 

exchangeable, 

carbonates

Step 2

40 mL of 0.50 M 

hydroxylamine 

hydrochloride plus 0.05 

M nitric acid, 16 h at 

room temperature

Amorphous iron 

and manganese 

oxides

Step 3

10 mL 8.8 M hydrogen 

peroxide at room 

temperature for 1 h plus 

1 h at 85 °C, reduce 

solution volume to near 

dryness; 50 mL 1 M 

ammonium acetate 

adjusted to pH 2 with 

nitric acid; 16 h at room 

temperature

Organic matter, 

sulfides

Notes: see Rauret et al. (1999) for detailed experimental procedures.
59

(e.g., Gruebel et al., 1988). On the other hand, a number 

of complicating factors can strongly direct the outcome of 

sequential extraction procedures and the interpretation of 

results with regard to potential metal mobility. The two most 

identifiable experimental problems with sequential extraction procedures are the non-selectivity of extractants and 

potential trace element redistribution among phases during 

extraction (e.g., Rendell et al., 1980; Rapin et al., 1986). 

For example, metals that are released during extraction 

can potentially re-adsorb to other surfaces present in the 

sample at the time of extraction, or the extracting solution 

itself can impact geochemical conditions that favor metal 

removal by re-adsorption or precipitation. As noted previously in Section IIIB.3.1, extraction results are operationally 

defined and they are influenced by factors such as choice 

of reagents, extraction time, and solid to solution ratios.

An example of an extraction artifact relating to arsenic is 

described in Wilkin and Ford (2002). Extraction of samples 

with dilute hydrochloric acid is commonly adopted to assess 

metal and metalloid partitioning to the acid-volatile sulfide 

(AVS) fraction of soils and sediments. This extraction test 

is also used to estimate potential metal toxicity of aquatic 

sediments (DiToro et al., 1990; Allen et al., 1993; USEPA, 

2000b). The method takes advantage of the comparatively 

high solubility of metal monosulfides like FeS, PbS, CdS, 

and ZnS at low pH. However, insoluble metal sulfides of 

Cu and Hg are not effectively dissolved in hydrochloric acid 

(e.g., Mikac et al., 2000). Furthermore, arsenic chemistry 

in sulfidic systems contrasts with that of divalent metals in 

that the solubility of arsenic sulfides like orpiment (As2

S3) 

increases with pH; solubility minima for orpiment are found 

at low pH. This means that at the low pH conditions typical 

of HCl extractions, precipitation of arsenic sulfide is favored. 

Two potential artifacts complicate the use of low pH extractions for As solid-phase partitioning in sediments containing 

acid-volatile sulfides. If orpiment is present in sediment, 

it will not be efficiently dissolved with hydrochloric acid. A 

more serious problem, however, relates to interpretations 

of arsenic extraction data. During acid-extraction, arsenic 

may be released from labile sediment components, i.e., 

loosely bound or sorbed sites. If AVS is present, arsenic 

sulfide is expected to precipitate at low pH and thereby significantly impact arsenic partitioning by transferring arsenic 

from a labile and potentially bioavailable fraction to what 

would be considered, using conventional interpretations, a 

refractory, bio-unavailable fraction (i.e., a fraction insoluble 

in hydrochloric acid). 

A second example of element redistribution during chemical extraction procedures relating to lead is described in 

Gerth (1990) and Ford et al. (1999). In these studies metal 

partitioning to iron hydroxide was evaluated using a 2 h 

extraction with ammonium oxalate. The procedure takes 

advantage of the rapid dissolution of ferrihydrite in ammonium oxalate relative to other more stable transformation 

products of ferrihydrite, such as goethite. However, an 

insoluble Pb-oxalate phase was identified as an end-product 

of the chemical extraction. The formation of this insoluble 

phase was clearly a consequence of the choice of reagents 

used in the extraction procedure. Again without knowledge 

of this artifact, interpretation of extraction results would 

have likely guided the conclusion that lead was immobile 

and non-bioavailable, which in fact may or may not be the 

case. The magnitude of extraction non-selectivity and/or 

extraction artifacts can be assessed through the addition 

of specific mineral components targeted by the individual 

extraction steps. These internal reference materials can 

be synthesized in a manner that includes the contaminant 

in a sorbed form. Synthesis methods for many of the 

solid phase components targeted by the various proposed 

extraction protocols are included in the cited references. 

The inclusion of this form of quality control into the overall 

extraction methodology will increase the level of confidence 

in these analytical data.

IIIB.2.5 Attenuation Capacity

The capacity of geological materials to attenuate contaminants has been a subject of extensive research. The 

static or batch adsorption method has often been used 

to assess the capacity of soils and aquifer materials to 

remove metal components from the aqueous phase. The 

ease of carrying out batch-adsorption tests clearly factors 

into the popularity of this testing strategy. In this method, 

aqueous solutions are mixed with a given mass of solid 

material for a set period of time. The aqueous solution is 

next separated from the solid adsorbent material, typically 

by filtration, and chemically analyzed to determine changes 

in solute concentrations. The concentration of a dissolved 

component before reaction with the adsorbent minus the 

concentration after reaction is used to calculate the mass 

of the component that has been removed by adsorption 

(or some other removal process). The approach is simple 

yet numerous experimental parameters impact the results 

of a given batch-adsorption test (USEPA, 1992; USEPA, 

1999; Jenne, 1998).

For inorganic species of concern, critical experimental 

parameters include contact time, solution pH, method of 

mixing, solid:solution ratio, and the concentration of other 

dissolved components in the solution (e.g., Barrow and 

Shaw, 1979; Roy et al., 1986). Solution composition is 

of critical importance since dissolved constituents in the 

background matrix of the ground water can influence contaminant partitioning either through competition for sorption 

sites or modification of the surface charge of the aquifer 

matrix. In addition, due to chemical interactions between 

the aquifer matrix and the solution in which it is suspended, 

it is important that solution matrix employed closely mimics 

the in-situ conditions from which the aquifer matrix was 

collected. For example, exposure of reduced sediments 

to a solution matrix containing molecular oxygen will likely 

cause significant changes in the mineralogical composition 

of the aquifer matrix during the time period of the batch 

test (USEPA, 2006). The results from a test performed 

in this manner provide no meaningful data relative to the 

assessment of in-situ sorption properties.

The methods used to prepare samples for use in laboratorybased studies, including batch-adsorption tests, can have a 

profound influence on test results. For example, oven-drying 
60

samples, although useful for homogenizing materials and 

for obtaining accurate dry weights, in most cases affects 

the chemical properties of sediments and may influence 

the results of batch-adsorption procedures. This affect will 

vary from material to material but may be the consequence 

of surface area changes that come about after drying, mineral transformations that occur at elevated temperatures or 

result from oxidation reactions. Consequently, oven-drying 

is generally not an advisable practice to accelerate material 

drying. Air-drying is more preferable although this method 

may take a longer period of time (several days). Air-drying 

can be considered to be a partial drying procedure to bring 

the moisture content of a material to near equilibrium with 

the atmosphere under which drying is taking place. The 

duration of air-drying should be kept to the minimum time 

possible and the progress should be tracked with methods 

such as weighing. Anaerobic soils provide a special case. 

Drying of such materials must be carried out in a anaerobic chamber or glove box. Drying of anaerobic materials 

can be accelerated by using a stream of dry, high-purity 

(oxygen-free) inert gas.

Batch-adsorption tests should be carried out under constant-temperature conditions (e.g., ±3 °C) and should attempt to mimic site chemical conditions. The influence of 

pH on batch-adsorption tests is extremely important and 

the effect of pH will vary depending on the nature of the 

adsorbent and the solute of interest. Measurements of 

the equilibrium pH of the soil-water suspension should be 

given along with adsorption results. For anaerobic systems, 

batch tests and pH measurements should be carried out 

in a glove box so that air-sensitive soil components do not 

oxidize and influence contaminant sorption (USEPA, 2006). 

Methods for ensuring proper mixing of solid and solution 

mixtures and selection of appropriate solid: solution ratios 

are discussed for example in USEPA (1992).

Results of batch-adsorption tests are conveniently analyzed using linear regression tools and adsorption isotherm 

equations that relate the amount of solute adsorbed to the 

equilibrium concentration of the solute. Two of the more 

frequently used adsorption models employ the Langmuir 

Equation (Section IIIB.3.2) and the Freundlich Equation 

(Section IIIB.3.3). The choice of one of these adsorption 

models or other possible adsorption equations will typically 

stem from the simplicity of the equation and from statistical 

reasoning using the regression coefficient. These equations 

can be used to quantitatively describe adsorption data and 

be “plugged” into reaction transport models used to develop 

site models (see Section ID).

Dynamic, continuous flow column experiments are a more 

detailed and desirable method for obtaining metal uptake 

and desorption potential. In column experiments, aquifer 

materials are packed into a column apparatus, typically a 

glass chromatography column equipped with Teflon endplate assemblies. Columns typically contain sampling 

ports at the influent and effluent ends and preferably 

along the length of the column. The sampling ports are 

designed to allow for water sampling along the center axis 

of the column. Representative solutions are then pumped 

through the column using for example a high-performance 

liquid chromatography pump at a flow rate selected to 

approximate an average seepage velocity expected in 

the field. Effluent solutions and sampling ports along the 

column are then monitored with respect to contaminant 

concentrations, geochemical parameters (e.g., pH, redox 

potential), and volume of solution eluted from the column. 

These data allow for the construction of contaminant 

breakthrough curves based on the reduced contaminant 

concentrations (Ceffluent/Cinput) and sample pore volume. The 

results of column tests can be augmented by the application 

of mineralogical characterization techniques to allow for the 

identification of reaction products that can lead to insight 

regarding uptake mechanisms and provide an improved 

basis for predicting long-term trends.

The results of static-batch and dynamic-column tests ultimately will be fed into a mathematical model (i.e., a surface 

complexation model, or Freundlich adsorption isotherm) to 

develop a quantitative description of contaminant sorption 

to and desorption from aquifer materials. Examples of this 

methodology are presented for example in Dunnivant et 

al. (1992) and Kent et al. (1995), studies that explore the 

transport of cadmium and chromium/selenium, respectively, 

in aquifer systems.

In context of using laboratory- or field-derived data as input 

to computer modeling codes, Bethke and Brady (2000) 

have recently pointed out potential problems with the “Kd

” 

or single parameter distribution coefficient approach. When 

sorption is suspected as being the dominant attenuation 

mechanism, these authors argue that the simple distribution coefficient will in many cases not adequately describe 

contaminant movement through aquifer systems, especially 

for the ionic species typical of inorganic contaminants of 

concern in this document. The distribution coefficient 

simply gives the ratio of a metal ion’s sorbed concentration 

(mol/g sediment) to its dissolved concentration (mol/cm3). 

Alternatively, Freundlich or Langmuir adsorption isotherms 

or surface complexation models are available to more 

accurately model field observations. The advantage of 

the isotherm approach is that adsorption trends can be 

tied specifically to materials collected from a specific site, 

yet the method is empirical and it must be acknowledged 

that extrapolation of results can not be made outside of 

measured ranges. The surface complexation model does 

indeed provide a more realistic description of ion adsorption from fundamental principles, yet available databases 

for surface complexation constants are limited. Almost all 

applications of surface complexation modeling efforts use 

hydrous ferric oxide as the dominant sorbing material, this 

may not be appropriate for all sites.

IIIB.3 Model Representations to Interpret 

Contaminant Sorption Observations

While thermodynamic equilibrium-based geochemical 

models are useful for providing boundary conditions for 

estimating the extent of contaminant partitioning to aquifer 

solid phases, the kinetics and reversibility of the sorption 
61

process are factors that warrant consideration on a sitespecific basis if monitored natural attenuation is proposed 

as a remedial strategy for inorganic contaminants. Another 

important consideration is that a purely thermodynamic 

treatment of partitioning is a purely macroscopic description 

and therefore, is not dependent on an atomic or a molecular 

scale model which may be employed to represent sorption 

reactions occurring within the aquifer (Sposito, 1981). 

There are two general approaches to modeling contaminant sorption behavior, namely empirical and mechanistic 

models. Empirical models provide a mathematical description of observed experimental data without necessarily 

invoking a theoretical basis or microscopic model for the 

observed relationship. Mechanistic models seek to describe a system based on thermodynamic principles. In 

theory, the mechanistic approach is desirable because it 

is more robust and widely applicable in that the effects of 

changes in ionic strength and pH within a system are fully 

integrated into the models. On the other hand, mechanistic 

models require a rather complete chemical and physical 

description of the system to be modeled and thus are still 

less commonly used than the empirical models. In general, 

mathematical descriptions of contaminant sorption to solid 

surfaces capture the electrostatic and/or chemical forces 

(or some combination thereof) that result in a net attraction 

of dissolved contaminants to the solid surface. Empirical 

models are derived to describe sorption trends irrespective 

of the specific mechanisms, e.g., electrostatic/chemical 

forces, involved in the partitioning process. In contrast, 

mechanistic models attempt to describe the relationship 

between properties of the solid surface and the net attraction/repulsion of the dissolved contaminant. 

IIIB.3.1 Distribution Coefficient/Partition 

Coefficient, Kd

The partition coefficient, Kd, is the simplest model for predicting sorption in soil systems. It is defined as the ratio of 

the quantity of adsorbate sorbed per unit mass of solid to 

the quantity of adsorbate in solution at equilibrium:

Kd = q/C

where q = concentration of adsorbate on the solid at 

equilibrium (μg/g) and C = total dissolved concentration 

remaining in solution (μg/ml).

This approach assumes that the system is reversible and 

that sorption is independent of the adsorbate concentration 

in the aqueous phase. Like all the empirical models the 

constant Kd

 value does not account for changing physical 

and chemical conditions in the soil. An extension of the 

constant Kd is the parametric Kd model where the dependence of Kd of a particular contaminant on various physical 

and chemical properties of the system is determined by 

stepwise linear regression analysis and polynomial expressions are developed that express Kd

 as a function of the 

relevant soil and aqueous conditions. The parametric Kd

approach is preferable in that it allows for the Kd value to 

vary dependant on prevalent conditions, however, it arguably involves as complete an analytical characterization of 

the system as the mechanistic models while still remaining 

an empirical approach.

The basic tenant of the constant Kd approach, i.e., that 

partitioning is a linear function of concentration, has been 

shown to be invalid in many instances. Adsorption isotherm 

models have been used by soil scientists to model situations were adsorption deviates from linearity. The two most 

commonly used models are the Langmuir and Freundlich 

adsorption isotherm models.

IIIB.3.2 The Langmuir Model

The Langmuir model was originally developed to describe 

the adsorption of gas molecules on a homogenous solid 

surface. The assumptions underlying the model are that 

every adsorption site is equivalent, and that the ability of 

a site to bind the adsorbate is independent of whether 

neighboring sites are occupied. The Langmuir equation 

is expressed as:

q

bKC

KC = 1+

Where q is the concentration of adsorbate on the solid and 

C is the concentration in solution, b is the maximum number 

of available sites for adsorption (assumes monolayer coverage) and K is a constant related to binding strength. The 

Langmuir equation can be rearranged to a linear form:

Let Kd = q/C

then Kd = bK – Kq

and a plot of Kd versus q should be linear with a slope of 

–K and intercept of b. 

Sposito (1984) reports that it is not uncommon for the 

relationship between Kd and q to be convex to the q axis 

rather than linear. This type of isotherm has been fit with 

a two site Langmuir isotherm:

q

b K C

K C

b K C

K C = +

+

+

1 1

1

2 2

2 1 1

The adherence of an adsorption isotherm to a two site 

Langmuir model has been interpreted as evidence for 

two discrete binding sites on the solid phase, however, no 

mechanistic interpretation can really be inferred from the 

goodness of fit of sorption data to these isotherms.

IIIB.3.3 The Freundlich Isotherm

The Freundlich isotherm has the form:

q AC = i

β

Where A and β are adjustable parameters and β can have 

a value between 0 and 1. Hence a plot of log q vs. log C 

should be a straight line with an intercept of log A and slope 

of β. Note that when β = 1, then the Freundlich equation 

reduces to the linear Kd

.

Sorption isotherms have been widely used to describe and 

predict adsorption of a contaminant in soil and sediment 

systems, however, as stated previously they are purely 
62

empirically derived relationships and the validity of the calculated parameters can only be expected to hold within the 

bounds of the experimental data used to create the initial 

isotherm. They also provide no mechanistic interpretation 

of the reactions governing the solubility of a contaminant 

in a system; they are insensitive to the mode of sorption 

(i.e. precipitation or adsorption). In view of these facts 

mechanistic models of the soil sediment system have been 

developed. The mechanistic approach uses thermodynamic 

relationships to model aqueous speciation and electric 

double layer theory to model changes in surface charge 

as a function of pH, electrolyte concentration and valence. 

Mechanistic models are potentially much more robust than 

adsorption isotherm approaches and, once the model has 

been established, potentially much more powerful in their 

predictive capabilities with respect to changing physical 

and chemical conditions within the system.

IIIB.3.4 Mechanistic Models for Predicting 

Sorption - Surface Complexation

A variety of mechanistic approaches have been applied to 

provide a molecular description of adsorption in soil-water 

systems. A number of approaches are based on a mathematical description of the distribution of ions in the vicinity 

of a charged surface, and the surface charge and potential in the interfacial region. These mathematical models, 

termed surface complexation models (SCMs), capture the 

influence of electrostatic forces between the solid surface 

and charged ions within solution as well as the influence 

of chemical interactions between these two entities that 

leads to the formation of coordinative bonds. The various 

models that are commonly employed to describe surface 

complexation reactions differ in their description of the 

distribution of charge at the solid-water interface and how 

ions are distributed in the aqueous layer that bathes the 

solid surface. Details of the various models are available 

from several sources and should be consulted by those 

unfamiliar with the implementation of surface complexation 

modeling to describe contaminant solid-solution speciation 

(e.g., Sposito, 1984; Davis and Kent, 1990; Stumm, 1992; 

Goldberg, 1995). A brief synopsis of the various models 

is provided in Table 3.4. In general, two different types of 

surface complexes are proposed in the various models: 1) 

inner-sphere surface complexes in which the contaminant 

forms a bond with a surface functional group, and 2) outersphere surface complexes where the contaminant partitions 

to the surface via electrostatic attraction (similar to the 

process of ion exchange).

A key difference between SCMs and empirical models is the 

employment of thermodynamic concepts to model chemical reactions occurring at the solid surface. By proposing 

specific chemical reactions for the partitioning of solution 

ions to the solid surface it becomes feasible to account for 

the influence of the bulk solution composition such as pH, 

which can exert an influence on both surface charge as well 

as solute speciation. In addition, the bulk water composition can influence contaminant sorption via the presence 

of ions in solution that compete for available surface sites. 

Examples of surface complexation reactions that may be 

postulated within an SCM are as follows:

=SH + Zn2+ ↔ =S-Zn+ + H+ 

(=SH represents functional group on FeS)

=SH + Mg2+ ↔ =S-Mg+ + H+

=SH + OH- ↔ =S- + H2O

=FeOH + Pb2+ ↔ =FeOPb+ + H+

(=FeOH represents functional group on FeOOH)

=FeOH + Ca2+ ↔ =FeOPb+ + H+

=FeOH + H+ ↔ =FeOH2

+

A mass action equation can then be written for each of 

these reactions (similar to solution complexation reactions) 

and solved to calculate conditional surface complexation 

constants. For SCMs that account for the influence of 

electrostatic interactions, the conditional surface complexation constant is typically modified via multiplication by 

an exponential electrostatic factor in order to calculate an 

instrinsic surface complexation constant. 

Table 3.4 Synopsis of the various surface complexation models (SCMs) commonly employed to 

describe solute partitioning to solid surfaces.

Surface 

Complexation 

Model

Description 

of Solute 

Partitioning 

to Solid 

Surface

Influence of 

Surface Charge

Constant 

Capacitance 

Model 

(CCM)

Inner-sphere 

complexes

Exponential 

electrostatic term 

that modifies 

value of surface 

complexation 

constant

Diffuse Layer 

Model 

(DLM)

Inner-sphere 

complexes

Exponential 

electrostatic term 

that modifies 

value of surface 

complexation 

constant

Triple Layer 

Model 

(TLM)

Inner- and 

outer-sphere 

complexes

Two exponential 

electrostatic terms 

that modify values 

of inner-sphere 

and outersphere surface 

complexation 

constants

Nonelectrostatic 

Model (NEM)

Inner-sphere 

complexes None
63

The reactions describing ion partitioning to a solid surface 

provide the physicochemical context for assessing changes 

in sorption that may accompany changes in ground-water 

chemistry. Several observations are evident from the example reactions shown above:

1) pH can influence the speciation of surface functional groups as well as the surface charge,

2) ion sorption can influence surface charge, and

3) major ions in solution (e.g., Mg2+ and Ca2+) can 

compete with the contaminants of concern (e.g., 

Zn2+ and Pb2+) for available sorption sites.

Thus, an SCM provides one with the ability to project how 

changes in solution chemistry can impact contaminant 

uptake; a significant limitation of empirical partitioning 

relationships. While this flexibility provides a powerful tool 

for examining the evolution of a contaminant plume, current 

implementations of most SCMs is limited by the heterogeneous nature of solid surfaces within an aquifer (types of 

sorption sites) and the ability to predict surface charging 

behavior. Recent efforts have been employed to develop 

nonelectrostatic surface complexation constants for sitespecific descriptions of subsurface contaminant transport 

(e.g., Kohler et al., 2004). This approach will most likely be 

successful for contaminants that predominantly form strong 

chemical bonds with available solid surfaces.

IIIB.3.5 Mineral Solubility

Characterization of the solid phase will, in most cases, be 

an integral component of monitored natural attenuation 

assessment and application. Some insight regarding the 

mineralogical composition of aquifer systems can be obtained by analyzing solution compositions. Investigations 

of mineral solubility in aquifer systems usually concentrate 

on the question: does a mineral control the concentration of 

a particular element, and, if so, what is the identity of the 

mineral? In order to answer these questions, geochemical 

models are frequently employed to calculate ion activities 

and mineral saturation indices. In aquifer systems, some 

chemical reactions are sufficiently fast that equilibrium 

relationships are immediately attained. For example, protonation/deprotonation reactions of acids and bases and 

ion pairing reactions are fast chemical reactions. Other 

reactions, in particular those reactions involving solids, 

can proceed so slowly that equilibrium is not attained even 

after decadal time periods. Yet, equilibrium is a practical 

reference point and equilibrium relationships are useful for 

predicting reactions that are likely or unlikely to occur.

Mineral solubility is influenced by the ionic strength of 

solution. Unless conditional equilibrium constants are employed, solution activity models are a necessary component 

of geochemical modeling efforts. Activity models account 

for the non-ideal behavior of solute ions in aqueous solutions. Non-ideal behavior is a consequence of electrostatic 

interactions between water molecules and charged solute 

ions. Methods for computing individual ion activity coefficients are presented in numerous textbooks (e.g., Stumm 

and Morgan, 1981; Nordstrom and Munoz, 1986). For 

most ground water studies of low ionic strength waters, the 

extended Debye-Hückel equation or the Davies equation 

will provide reasonable activity models. In these equations, 

activity coefficients are estimated based on input values for 

solution ionic strength and temperature. For concentrated 

waters that are high in total dissolved solids, virial methods 

such as the Pitzer model are available for estimating ion 

activity coefficients (Langmuir, 1997).

The unit of concentration for dissolved species most frequently used for aqueous solutions is molality, mi

 (mol/kg). 

Analytical concentrations are often expressed in mass 

based units, e.g., ppm (parts per million). A useful conversion is:

Conc. in ppm = Conc. in mol/kg ×

formula weight in g/mol × 1000

The effective concentration or activity of a dissolved species, ai

, is given by:

ai

 = (γi mi)/(γi

0 mi0

)

Where γi is the ion-specific activity coefficient and (γi

0 mi0

) 

refers to the standard state, which for aqueous solutions 

is typically chosen as an ideal, 1 molal solution, i.e., both 

γi

0

 and mi

0

 are equal to 1.

In order to evaluate whether a ground water is oversaturated, undersaturated, or at equilibrium with a particular 

phase, geochemical speciation models are of practical 

use. As an example, consider the solubility expression for 

cadmium sulfide (greenockite):

CdS(s) + H+ = Cd2+ + HSThe mass-action expression that applies to the equilibrium 

is:

K

a a

a a r

Cd HS

CdS H

= = + −

+

− 2 14 4 10 .

Ground water may or may not be at saturation with respect 

to greenockite, depending on whether the phase is indeed 

present, available surface area, residence time of water, 

and kinetic factors that may impede dissolution and/or 

precipitation. If the water is at equilibrium, then the ion 

activity product, Q, should be the same as the equilibrium 

constant, i.e., 

Q

a a

a

K Cd HS

H

= = =r

+ −

+

− 2 14 4 10 .

where the activity of CdS is taken to be unity. Calculation 

of the saturation index (SI) for a water can then be used to 

determine if the solution is undersaturated, at equilibrium, 

or oversaturated with respect to precipitation (see Section 

IIB.2.1, Table 2.3).

It is important to point out that solubility products for precipitates, such as cadmium sulfide, depend on whether 

the solid is freshly precipitated (typically disordered) or 

crystalline (well ordered). At 25 °C, Daskalakis and Helz 

(1992) report a solubility product of 10-14.36 for crystalline 

cadmium sulfide (greenockite). Wang and Tessier (1999) 
64

give a value of a freshly prepared CdS precipitate of 10-14.15. 

Hence, freshly precipitated cadmium sulfide is metastable 

with respect to greenockite since:

CdS(disordered) = CdS(greenockite) Kr

 = 100.21

and

∆Gr

0

 = -αRTlog K = -1.20 kJ/mol

If it is possible to calculate activities of free ions from a given 

water analysis, then it is a relatively simple matter to calculate the degree saturation of a large number of possible 

mineral phases, keeping in mind that temperature and ionic 

strength are needed to correct the ion activity product.

In the example above, pH and the molal concentrations 

and ion activity coefficients for Cd2+ and HS-

 are needed 

in order to compute the ion activity product. If complexes 

of cadmium are present, then mCd

2+ is not the same as the 

total concentration of dissolved cadmium. If we consider 

only mononuclear bisulfide complexes,

m m m m m Cd total Cd CdHS Cd HS Cd HS , = + + + + + ( ) ( ) 2 − 2

0 3

Depending on the solution composition, cadmium complexes with chloride, bicarbonate, hydroxide, or sulfate may 

be important in addition to bisulfide complexes. Similarly, 

mHS- is not the same as the total concentration or analytical 

concentration of dissolved sulfide. Both protonated and 

deprotonated forms of sulfide and complexed forms of 

sulfide may be present. Speciation calculations, therefore, 

are critically dependent on the completeness and quality of 

thermodynamic constants in the thermodynamic database 

used for geochemical modeling (see Section ID).

In cases where mineral solubility is controlled in part by 

redox conditions, calculation of the saturation index is not 

straightforward. For example, when an aqueous solution 

has attained saturation with respect to a mineral such as 

the iron oxide magnetite (Fe3O4), then the reaction

Fe3O4 + 8H+ = Fe2+ + 2Fe3+ + 4H2O

is at reversible equilibrium when

a a a

a

K Fe Fe H O

H

r

2 3

2

2

4

8

4 2 10 + +

+

= = .

Magnetite solubility depends therefore on pH and redox. It 

follows that to calculate the saturation index for magnetite, 

activities of both the ferrous and ferric ions are needed input 

variables. Accurate measurement of ferric iron concentration at circumneutral pH is a notoriously difficult task. There 

are several ways to approach this problem:

• Measure separately both Fe(II) and Fe(III) in a water 

sample and estimate activity coefficients in order to 

calculate the activities of Fe2+ and Fe3+

• Use Eh as a master redox variable to fix Fe(II):Fe(III) 

ratios

• Measure Fe(II), calculate Fe(III) by assuming control 

by the solubility of Fe2O3, FeOOH, or Fe(OH)3

• Use some other measured redox pair to fix Fe(II)/

Fe(III) 

There are methods available to determine separately the 

concentration of ferrous iron and ferric iron (e.g., To et al., 

1999); however, these methods are most effective for use in 

low pH waters, such as aquatic systems impacted by mine 

wastes. At near-neutral pH, the concentration of Fe(III) is 

typically very low so that analytical detection limits constrain 

the use of direct measurements.

The second method is to use the measured Eh of a water 

to estimate the Fe(II)/Fe(III) ratio in solution. Applying the 

Nernst equation to the iron(II/III) couple, we have

E E RT

nF

a

a

a i

i Fe

Fe

= − = − +

+

0 2

3

ln . . log Π 0 770 0 0592 ν

Using this approach, Fe(II)/Fe(III) can be calculated from 

a potential measurement using a platinum electrode. Nordstrom et al. (1979) conclude that this approach is most 

effective at pH <6 and at total iron concentrations exceeding 

10-6 molal. A third method is to assume that the solubility of a ferric-bearing mineral controls the activity of Fe3+. 

For example, consider the solubility expression of hydrous 

ferric oxide:

Fe(OH)3 + 3H+ = Fe3+ + 3H2O Kr = 10-36.6

(Langmuir, 1996)

So that,

log log log a K a pH pH . Fe3 H O2 3 3 36 6 3 + = − − ≅ − −

In this example, the activity of Fe3+ can be estimated based 

on a measurement of pH. In the last example, a separate 

redox pair, e.g., As(III)/As(V), can be used to fix the Fe(II)/

Fe(III) ratio in solution by applying the Nernst equation. The 

correctness of any one of these approaches relies on how 

closely the assumptions are obeyed for a given system.

IIIB.3.5.1 Coprecipitation Reactions

The concept of a solid solution implies an isomorphic substitution, for example, regular substitution of Mn for Ca in 

a carbonate mineral. Minerals formed in the environment 

often contain substitutional impurities and coprecipitation 

is likely to be a primary natural attenuation mechanism. 

Thermodynamic models of solid solutions indicate that the 

solubility of a component becomes greatly reduced as that 

component becomes a constituent in a solid phase. In other 

words, solid solutions in binary systems, for example, are 

less soluble (more stable) than pure end-member compositions. A considerable effort has gone into understanding the 

thermodynamic and kinetic factors that control the formation 

of solid solutions from aqueous solutions (e.g., Lippmann, 

1977; Busenberg and Plummer, 1989; Glynn et al., 1990; 

Glynn and Reardon, 1990).

Laboratory and field studies indicate that cadmium is effectively removed by calcerous materials (e.g., Dudley et al., 

1988; Davis et al., 1987). The solid solution (Ca,Cd)CO3

 is 

complete and the endmember solubility products differ by 

three orders of magnitude at room temperature:

CaCO3 = Ca2+ + CO3

2- log KCaCO3

= -8.3

and
65

CdCO3 = Cd2+ + CO3

2- log KCdCO3

= -11.3

Combining these solubility expressions we obtain an exchange reaction:

CaCO3 + Cd2+ = CdCO3 + Ca2+

The equilibrium constant for this type of reaction is often 

referred to as a distribution constant, D, and is given by the 

quotient of the solubility products of CaCO3 and CdCO3:

D

a

a

a

a

K

K

CdCO

CaCO

Cd

Ca

CaCO

CdCO

= 









 = = +

+

−

3

3

2

2

1

3

3

3  10

The activity ratio of the solids by definition can be replaced 

by the ratio of the mole fractions multiplied by activity coefficients:

D

X

X

a

a

CdCO

CaCO

CdCO

CaCO

Ca

Cd

=





































+

+

3

3

3

3

2

2

  γ

γ

Therefore, the amount of substitution of Cd into calcium 

carbonate is a function of the solubility product ratio of 

CaCO3

 to CdCO3, the solution activity ratio of Cd2+ to Ca2+, 

and a term that represents activity coefficients of the solid 

solution components:

X

X

K

K

a

a

CdCO

CaCO

CdCO

CdCO

Cd

Ca

CdCO

Cd

3

3

3

3

2

2

3 









 = 











+

+

  γ

γ CO3













In general practice the first two terms, solubility product 

ratio and the solution activity ratio, are straightforward to 

obtain. Activity coefficients in the solid phase in most cases 

differ markedly from 1. In principle, the activity term can 

be determined experimentally by measuring the mole fraction ratio over a range of solution activity ratios at constant 

temperature.

An analysis of solution concentration data alone will 

generally not be adequate to confirm any precipitation or 

coprecipitation mechanism of attenuation. Multiple lines of 

evidence, including solution and solid-phase data, which 

supports a specific natural attenuation mechanism will in 

most cases be necessary to pursue monitored natural attenuation as a partial or sole cleanup remedy.

IIIB.3.5.2 Thermodynamic Data

Site characterization and remedial investigations that 

involve contaminant transport in the environment are usually accomplished, in part, with geochemical modeling. 

Consequently, thermodynamic properties are essential 

components to geochemical modeling efforts. Normally, the 

thermodynamic constants used in modeling exercises are 

taken from published compilations of such properties (e.g., 

Wagman et al., 1982). In some cases these compilations 

contain minimal documentation as to the source or quality of compiled data. In other cases, compilations contain 

outdated information. Users of thermodynamic databases 

can easily be misled into believing that modern databases 

are up to date in terms of data quality and completeness.

As an example to illustrate problems noted above, a specific example of orpiment (As2

S3(s)) is discussed. Eary 

(1992) listed a number of values of the Gibbs free energy 

of formation, ∆Gf

0

, of orpiment, each from a separate literature source. The values tabulated by Eary (1992) range 

from –168.8 kJ/mol to –90.7 kJ/mol. This wide range in 

Gibbs free energy values equates to equilibrium constants 

describing orpiment dissolution and precipitation that can 

span almost 7 orders of magnitude. For example, consider 

the solubility expression for orpiment to form arsenite in 

solution:

0.5As2

S3 + 3H2O = As(OH)3

0

 + 1.5HS- + H+

An equilibrium constant for this reaction can be computed 

from the ∆Gf

0

 values for orpiment given above and fixed 

values of ∆Gf

0

 for H2O(l), As(OH)3

0

, and HS- of –639.8 kJ/

mol, -237.18 kJ/mol, and 12.05 kJ/mol, respectively. The 

more negative ∆Gf

0

 value for orpiment gives a log K value 

for the above reaction of –30.5, while the greater value gives 

a log K value of –23.7. The pH dependent trend of orpiment solubility resulting from these differing log K values is 

shown in Figure 3.5. The solubility diagram is constructed at 

ΣH2S concentration of 10-5.5 m, so that thioarsenite species 

have a negligible contribution to orpiment solubility (Wilkin 

et al., 2003). It is clear that depending on the log K value 

employed in reaction modeling, solutions will have vastly 

different saturation indices and one would reach different 

conclusions about possible attenuation mechanisms for 

arsenic.

Figure 3.5 pH-dependent solubility trend of orpiment

predicted using two different Gibbs free

energy of formation values (see text). The

model curves correspond to a constant

ΣH2S concentration of 10-5.5 m. The data

points are measured solubility data taken

from Webster (1990). The experimental

data are consistent with a ∆Gf

0 for orpiment of approximately –90 kJ/mol, yet most

thermodynamic compilations adopt a value

of –168.8 kJ/mol. Application of the lower

value will vastly over predict the stability of

orpiment.
66

Which value of ∆Gf

0

 for orpiment is more appropriate? 

The most recent solubility and calorimetric studies seem 

to support a ∆Gf

0

 formation value close to –90 kJ/mol for 

crystalline orpiment (Webster, 1990; Eary, 1992; Johnson 

et al., 1980). The key point is that users of geochemical 

models should be familiar with the sources of thermodynamic data used and attempt to evaluate whether constants 

in thermodynamic compilations are in reasonable agreement with recent experimental evidence or whether there 

may be discrepancies that could impact how model results 

are interpreted.

IIIC. Characterization of System Redox and 

Underlying Microbial Processes

Oxidation-reduction processes affect the chemical composition of ground water and impact the aqueous and solid 

phase chemical speciation of inorganic contaminants. Thus, 

characterizing the redox status of ground water systems 

will likely play an important role in understanding controls 

on contaminant attenuation. As stated previously in Section IIC, the subsurface microbiology within a plume will, 

in part, influence the predominant redox characteristics of 

the system. Microbial processes may influence the redox 

chemistry of important components within the aquifer, 

such as iron and sulfur, participating as reactants within 

attenuation reactions that result in contaminant immobilization or degradation. Thus, it is recommended that site 

characterization include both assessment of the prevailing 

subsurface redox chemistry impacting contaminant transport, as well as more explicit analysis of specific microbial 

processes for sites at which the plume conditions support 

microbial activity that differs from ambient conditions. For 

example, a more detailed microbial community analysis 

may be warranted in order to improve the reliability of assessing attenuation capacity for sites where contaminant 

source characteristics govern aquifer geochemistry within 

the down gradient plume. The following discussion provides 

approaches and tools available to characterize the overall 

redox status of site ground water, as well as the specific 

microbial communities that support the observed redox 

chemistry within the plume.

IIIC.1 Process Identification

Identification of a redox-mediated attenuation process starts 

with knowledge of the aqueous and solid-phase species 

present. Although analysis of the contaminant concentration in ground water at various distances from the source 

can indicate attenuation, inclusion of data for pH, EH, major 

solutes, and redox-sensitive species such as Fe2+, O2, and 

H2S is needed to demonstrate that redox processes may 

be involved. 

Further evidence for redox-mediated processes may be 

derived from determination of the mineral species present. 

The absence of Fe and Mn oxides and the presence of 

Fe sulfides, for example, suggest that the aquifer is predominantly reducing, whereas the inverse of this situation 

suggests an oxidizing aquifer. A variety of techniques can 

be used for mineral identification and is detailed elsewhere 

in this document (Section IIIB.2). A major consideration 

related to mineralogical characterization relates to the 

preservation of original mineralogy in aquifer materials by 

preventing contact with oxygen during sample collection and 

removal of pore water. For example, iron-bearing minerals 

may exist in a reduced state within the saturated zone. Yet 

such minerals that are stable in reducing environments 

are subject to significant alteration upon exposure to oxygen. Solid phase structural and chemical transformations 

are commonly mediated or facilitated by the pore water. 

Thus, removal of pore water may act to retard or impede 

transformation. In order to preserve redox characteristics 

of samples collected in the field, cores materials should 

be immediately capped and frozen. Sample freezing can 

be accomplished either by submersing in liquid nitrogen or 

placement in a portable freezer located in the field. Following transport to a laboratory setting, frozen materials should 

be thawed under an oxygen-free or inert atmosphere, 

e.g., within an anaerobic glove box. Extended periods of 

storage should be avoided, since sample mineralogy will 

alter with time during approach to a new equilibrium state. 

Procedures for the collection and processing of aquifer 

solids to minimize alterations to mineralogy are outlined 

in several EPA documents (USEPA, 2002; USEPA, 2005b; 

USEPA, 2006). 

In some situations, intense local microbial activity may be 

entirely responsible for the redox status of the aquifer. The 

nature of the active microbial population (e.g., iron-reducing, 

sulfate-reducing, or sulfur-oxidizing bacteria) can often be 

inferred from geochemical data. Thus, trends in the concentration of organic substrates (soluble organic C) and their 

metabolites (e.g., H2

, H2S, CH4, CO2, NO2

-

, HS-, Fe2+) can 

indicate whether and which microorganisms are active in a 

particular subsurface region. In some instances, direct and 

specific determination of microbial population by culturing or 

genetic analysis (e.g., messenger ribonucleic acid profiles) 

of aquifer solids extracts may be warranted.

Although the best support for a particular redox process 

comes from direct examination of reactants and products 

at the site, a strong case for the process can be made with 

laboratory tests using core materials taken from the site. 

These tests involve batch or column studies in which the 

geochemical inputs and conditions are comparable to those 

at the actual site, and evidence for the process is obtained 

from analysis of the solution phase (supernate or effluent) 

during the test and of the solid phases at its conclusion. 

Attenuation obviously must be demonstrated through the 

solution-phase data, and direct identification of the attenuated form of the contaminant provided from the solid-phase 

data. As described below, this same general approach with 

suitable modifications can be used to determine capacity, 

kinetics, and stability.

IIIC.1.1 Redox Measurements

Measurement of redox parameters in ground water is inherently challenging due to the fact that a steep redox gradient 

is often present between the sampling location (subsurface) 

and the location where the particular measurement is made 
67

(surface). Consequently, special care must be taken to 

preserve the redox integrity of ground-water samples. In 

some cases water samples can be preserved by using inert 

gases or by acidification because oxidation rates generally 

decrease substantially with decreasing pH. In other cases 

where sample preservation is not possible or practical, it is 

recommended that analyses be carried out in the field at the 

time of sample collection (e.g., for dissolved oxygen and Eh). 

Summaries of methods used to determine redox status can 

be found in, for example, USEPA (2002), Baedecker and 

Cozzarelli (1992), and Walton-Day et al. (1990). Table 3.5 

provides a list of parameters that can be used to assess 

the redox status of ground water systems.

Other laboratory approaches include selective dissolution 

of reactive fractions (see Section IIIB.2.4.1). For example, 

reactive Fe2+ (i.e., that which is sorbed as well as that present in sulfides carbonates, hydroxides, and green rusts) 

can be estimated by extraction with 0.5 M HCl followed by 

complexation with ferrozine (Roden and Zachara, 1996; 

Amonette et al., 2000). From the oxidizing side, Mn(III,IV) 

oxides and poorly ordered Fe(III) (hydr)oxides can be determined by extraction with hydroxylamine solutions (Chao, 

1972; Chao and Zhou, 1983; Ross, 1985).

The direct reactions and laboratory tests, however, measure 

only the redox-buffering capacity at a single point in time. 

Microbial activity, through conversion of organic C, can 

create and replenish the reductive capacity of a site. Thus, 

more accurate measurements of reductive capacity may 

necessitate consideration of inputs of dissolved constituents 

from source areas or up gradient portions of the aquifer 

such as dissolved organic C, as well as terminal electron 

acceptors such as oxygen, nitrate, and sulfate, in order to 

assess ability of the microbial community to maintain or 

degrade the redox capacity of a site.

IIIC.2 Capacity

Once the operative redox-mediated attenuation process has 

been identified, an assessment of its capacity to attenuate 

the contaminant is needed. The primary soluble electron 

donor in most ground waters is dissolved organic carbon, 

although, under some circumstances, sulfide (H2

S, and HS-) 

Table 3.5 Ground-water redox parameters and measurement approaches.

Parameter Measurement Approach

Oxidation-reduction 

potential (ORP)

Combination platinum electrode with Ag/AgCl reference electrode; KCl filling solution. 

Electrode performance is determined using reference solutions (e.g., Zobell’s solution, 

hydroquinone).

Dissolved oxygen Membrane-covered electrodes; colorimetric tests, modified Winkler titration. Electrode 

performance is determined using air-saturated water and sodium sulfite solutions.

Dissolved hydrogen Sample collection in glass vessel and analysis by gas chromatography/reduced gas 

analyzer.

Iron speciation

Ferrous iron determined by colorimetric analysis (e.g., ferrozine, 1,10-phenanthroline). 

Ferric iron determined by adding reducing agent and measuring ferrous iron, and/or 

by determination of total iron and subtracting Fe(II). Measurement made in the field or 

preservation required.

Sulfur speciation

Sulfate and other sulfoxyanions (sulfite, thiosulfate) are typically determined by ion 

chromatography or capillary electrophoresis. Sulfide can be determined by colorimetric, 

gravimetric, coulometric, or voltametric methods. Measurement made in the field or 

preservation required.

Nitrogen speciation Colorimetric, chromatographic, and potentiometric methods available for nitrate, nitrite, 

and ammonium. Preservation of sample is recommended.

Arsenic speciation Anion exchange, chromatographic, and hydride generation methods are available for arsenite, arsenate, and several organic forms of arsenic. Preservation is recommended.

Chromium speciation Colorimetric, exchange, and voltametric methods are available for determination of Cr(VI) 

and Cr(III).

Selenium speciation Selenium oxyanions can be determined by ion chromatography or capillary 

electrophoresis. Preservation of sample is recommended.
68

and Fe2+ species can dominate. Important electron acceptors include oxygen, nitrate, nitrite, and sulfate. Thus, in 

addition to the contaminant, ground-water analyses should 

include measurements of these constituents. Equally 

important is a hydrological assessment of ground water 

flow rates, which, when combined with the concentration 

data, allows an estimate of the average influx of oxidants 

or reductants to the site. In addition, quantification of the 

accessible redox-buffering capacity of the aquifer solids 

may be important, since these solids may represent a 

significant fraction of the capacity of the aquifer. Methods for characterizing the oxidation capacity and reducing 

capacity of aquifer solids are listed in Table 3.6. These 

methods represent examples of approaches that have been 

tested and documented in the literature. For site-specific 

applications of these methods, some method development 

or modification may be required to obtain optimal results 

(USEPA, 2002).

Laboratory tests can be performed in which a known oxidant 

[e.g., O2(aq) or Cr(VI)] or reductant [e.g., H2S(aq)] is reacted 

with the aquifer solids under controlled environmentally 

relevant conditions (Fruchter et al., 1996; Istok et al., 1999). 

The quantity of this reagent consumed by reaction is then 

expressed in terms of the mass or volume of the aquifer 

solids. Rough estimates of maximum reductive capacity 

present in a soil can also be obtained from digestions using 

acidic Cr(VI) solutions. Thus, adaptations of Cr(VI) titration 

methods for organic C in soils (e.g., Nelson and Sommers, 

1996) are suitable as they include the contributions of Fe(II) 

and sulfides to overall reductive capacity. Other laboratory approaches include selective dissolution of reactive 

fractions (see Section IIIB.2.4.1) with the assignment of 

a specific mass-based oxidation/reduction capacity to the 

phase quantified by extraction. For example, reactive Fe2+

(i.e., that which is sorbed as well as that present in sulfides 

carbonates, hydroxides, and green rusts) can be estimated 

by extraction with 0.5 M HCl followed by complexation 

with ferrozine (Roden and Zachara, 1996; Amonette et 

al., 2000). From the oxidizing side, Mn(III,IV) oxides and 

poorly ordered Fe(III) (hydr)oxides can be determined by 

extraction with hydroxylamine solutions (Chao, 1972; Chao 

and Zhou, 1983; Ross, 1985).

The direct reactions and laboratory tests, however, measure only the redox-buffering capacity at a single point in 

time. Microbial activity, through conversion of organic C, 

can create and replenish the reductive capacity of a site. 

Thus, more accurate measurements of reductive capacity 

Table 3.6 Methods that may be employed for estimating the oxidation and reduction capacity for solid materials 

(from USEPA, 2002).

Method Source Comments

Cr(II) (oxidation capacity)

(Barcelona and Holm, 1991a)

(Barcelona and Holm, 1991b)

(Barcelona and Holm, 1992)

Most aggressive but oxygen-free atmosphere 

recommended; high estimate 

Digestion with Ti(III)-EDTA 

(oxidation capacity) (Ryan and Gschwend, 1991) Developed for extraction of Fe oxides; applicability limited to iron oxide dominated sediments

Titration/digestion with 

dithionite solution 

(oxidation capacity)

(Loeppert and Inskeep, 1996)

(Williams et al., 2000)

Valid for only specific remedial technology; targets 

Fe oxides 

Chemical Oxygen Demand 

by digestion with acid 

dichromate 

(reduction capacity)

(USEPA, 1979)

(Barcelona and Holm, 1991a)

(Barcelona and Holm, 1991b)

(Barcelona and Holm, 1992)

Precipitate coatings if pH not buffered; high 

estimate 

Digestion in hydrogen 

peroxide solution 

(reduction capacity)

(Nelson and Sommers, 1996) Developed for quantifying organic matter content

Dissolved oxygen 

consumption in 

air-saturated water 

(reduction capacity)

(Williams et al., 2000)

Dynamic column test with mathematical simulation; 

test design must minimize gas diffusion from 

external sources; time-consuming, but realistic
69

may necessitate consideration of inputs such as dissolved 

organic C, as well as terminal electron acceptors such as 

oxygen, nitrate, and sulfate, in order to assess ability of 

the microbial community to maintain or degrade the redox 

capacity of a site. For this analysis, it may be necessary 

to evaluate microbial response to dissolved constituent 

inputs through sampling of aquifer solids for the purpose 

of conducting microcosm studies to examine microbial activity and/or contaminant attenuation. Recommendations 

for the design and implementation of microcosm studies 

are provided in USEPA (1998). Evaluation of microcosm 

response to variations in electron donor/acceptor concentrations in solution, relative to contaminant attenuation, 

provides means for directly assessing the limits in these 

reactant concentrations under which attenuation remains 

viable. For situations in which contaminant degradation is 

the primary attenuation process, microcosm measurements 

will likely include determination of trends in contaminant 

loss as well as the increase of degradation products. For 

situations in which contaminant immobilization dominates, 

it is recommended that microcosm characterization include 

determination of the quantity and solid phase speciation 

of contaminant sorption. The degree to which microcosm 

studies replicate subsurface conditions within the plume 

may be assessed through comparison of similarities (or 

lack thereof) between measured aqueous and solid phase 

chemical parameters based on measurement of aquifer 

microcosm properties (e.g., water chemistry, mineralogy).

An additional line of evidence to support capacity assessments includes development of a reaction or reactive transport model that incorporates quantitative description of the 

processes that control contaminant attenuation. Requirements for model construction and parameter inputs have 

previously been specified in Section ID. The utility of this 

type of analysis is the ability to quickly assess a range of 

ground-water conditions that may influence the efficiency 

of modeled attenuation reactions. However, as previously 

noted, the degree of uncertainty in model predictions will be 

constrained by the accuracy of parameter inputs to represent aquifer conditions within the plume. Thus, verification 

of model performance is warranted to demonstrate the 

ability of the model to reproduce measured ground-water 

conditions prior conducted model tests of aquifer capacity 

to support attenuation.

IIIC.3 Stability

If a redox-mediated attenuation mechanism has been identified and a reasonable estimate of the capacity of the aquifer 

to attenuate the contaminant has been made, evaluation 

of the stability of immobilized contaminants is needed to 

assess the potential for contaminant remobilization due to 

anticipated changes in ground-water chemistry. This component of the site characterization effort may include direct 

measurements of contaminant stability via laboratory- or 

field-based evaluations, which could be supplemented with 

implementation of reaction or reactive-transport models 

that explicitly consider the solid-phase speciation of the 

contaminant. The ultimate goal of this effort is to gauge the 

response of the aquifer, from the perspective of contaminant remobilization, to changes in aquifer redox status that 

may be driven by future increases in the influx of dissolved 

components such as oxygen or the cessation of microbial 

processes that accompany decreased influx of degradable 

organic contaminants.

Contaminant stability may be estimated through laboratory tests constructed using aquifer solids collected from 

within the zone of contaminant attenuation. Controlled 

tests could be then be devised that evaluate contaminant 

response to changes in specific ground-water parameters 

that may result in release of the contaminant from aquifer 

solids. For sites in which the ground-water chemistry differs 

significantly within the plume compared to up gradient or 

ambient conditions, this may entail exposure of the aquifer 

solids to ambient ground-water samples. Alternative approaches may include systematic variation of one or more 

parameters identified as being critical to the stability of the 

form in which the contaminant is immobilized. For example, 

this may involve systematic variation in ground-water 

sample pH over a range that captures current conditions 

as well as anticipated conditions that may be reflected by 

the pH measured in background wells installed within the 

aquifer. Other parameters that might be assessed include 

those that reflect conditions that might develop as a result 

of potential land-use changes, including influxes of dissolved constituents that might compete for adsorption sites 

and/or may form soluble complexes with the contaminant. 

An alternative approach to assessing the stability of an 

immobilized contaminant may include the implementation 

of in-situ studies using single-well push-pull tests (Istok et 

al.,1997; Haggerty et al., 1998; Senko et al., 2002). These 

tests can similarly be devised to evaluate contaminant 

response to changes in ground-water chemistry through 

manipulation of specific parameters via mixing of synthetic 

solutions with ground water retrieved from the well and 

subsequently re-injected into the aquifer. The re-injected 

water is then allowed to react with the aquifer solids for 

short periods of time (days to weeks), and then several pore 

volumes are withdrawn and analyzed to assess the fate of 

the contaminant and/or analyze potential by-products that 

result during a re-mobilization reaction. Comparison of 

reagent concentrations with those of a non-reactive tracer 

injected with the reagent can be use to evaluate the degree 

of mixing between water sources with the reaction zone 

and to assess the overall rate of reaction. 

As noted previously, modeling studies may provide a 

supplementary line of evidence to assess the sensitivity of 

the immobilized contaminant to changes in ground-water 

chemistry. This type of analysis provides an indirect means 

to gauge contaminant response, as well as a means to assess the impact of other ground-water characteristics that 

may not be practically assessed via direct measurements. 

However, as previously noted, the degree of uncertainty 

in model predictions will be constrained by the accuracy 

of reaction expressions and parameter inputs to represent 

aquifer conditions within the plume. Ultimately, these tests 

may require more explicit analysis of the microbial community that supports existing conditions or mediates future 

changes to redox conditions within the boundary of the 
70

plume. Approaches to identify and quantify active microbial 

communities within the subsurface are discussed below.

IIIC.4 Microbial Community Characterization

Microbiological evidence to support the natural attenuation of inorganic contaminants as a remedial alternative 

in ground water involves the characterization of microbial 

population size, diversity, composition, physiological and 

genetic/phylogenic traits. This section addresses techniques used for the characterization of subsurface microbial 

communities. In addition, some microbiological sampling 

practices for the assessment of ground water are discussed 

and emerging methodologies are identified.

IIIC.4.1 Standard and Emerging Techniques 

Analyses focused on the composition and diversity of 

bacterial community structures cannot rely on traditional 

microbiological procedures alone. This is especially critical 

in subsurface ecological systems because the vast majority 

of the microbial communities that reside in that environment 

have not been cultivated using culture-dependent methods 

(Amann et al., 1995). Although the importance of emerging 

molecular approaches in subsurface microbiology will be 

stressed later in this section, a successful program will likely 

include a combination of culture-independent methods as 

well as traditional cultivation strategies. Often, molecular 

tools are applied to pure culture isolates harvested from 

defined media. In other words, traditional selective and 

enrichment techniques are used to develop specific microbial communities to be further characterized with molecular 

monitoring (i.e., genes coding for 16S rRNAs [16S rDNA] 

to identify potential gene expressions). Since standard 

microbiological techniques are readily available in most 

environmental laboratories, only a brief discussion of them 

is provided here (see also Table 3.7).

• Most probable-number (MPN) technique is a standard 

methodology used to estimate the number of specific 

physiological types of bacteria. Usually, a modified 

basal medium is amended with a carbon substrate and 

electron acceptors for total heterotrophic aerobes or 

anaerobes, denitrifying, iron-reducing, sulfate-reducing, and methanogenic bacteria (Fedorak et al., 1987; 

Lovley, 1991; Mahne and Tiedje, 1995; Chapelle et 

al., 2002; Tanner, 1989). A three-tube dilution series 

is often used to provide a 10-fold dilution. The same 

regiment is applied to semi-solid media to provide plate 

counts.

• Acridine orange direct count (AODC) is the most versatile technique for yielding a count of the total intact 

cells without differentiation for viability (Ghiorse and 

Balkwill, 1983). Differential staining for live vs. dead 

can be used via the Bac-lightTM method (Loyd and 

Anthony, 1995).

• Phospholipid ester-linked fatty acids (PLFA) is a popular assay used to identifty “biomarkers” to provide a 

quantitative insight into three important attributes of 

microbial communities including viable biomass, community structure, and metabolic activity (Lehman et al., 

1995). An estimation of non-viable populations can be 

accompanied through the measurement of diglyceride 

fatty acids (DGFA). Both assays are independent of 

the bias inherent in classical culturing techniques providing a more accurate estimation of in-situ microbial 

populations. The lipid biomarker analysis is, however, 

incapable of identifying every microbial species in an 

environmental sample because many species contain 

over-lapping PFLA. 

• Community-level physiological profile (CLPP) can be 

carried out using Biolog-GN plates (Biolog, Inc., Hayward, CA). This “phenotypic fingerprinting” assay is 

useful for screening bacterial isolates and consortia to 

establish correlation between their activity and composition.

• rRNAa can be used for the determination of microbial 

biomass (Loyd and Anthony, 1995). 

Table 3.7 Standard and emerging techniques for microbial community characterization.

Method Type of Information

Most probable-number 

(MPN) Enumeration, Cultural 

Acridine orange direct 

counts (AODC)

Enumeration, 

Morphological Cultural

Phospholipid ester-linked 

fatty acids (PLFA)

Enumeration, 

Biochemical

Community-level 

physiological profile (CLPP) Physiological

Microcosms are routinely prepared using subsurface 

cores and ground-water samples for the characterization of microbial communities. Recently, there has been 

an increased interest in the use microcosms to perform 

molecular community fingerprinting such as denaturing 

gradient gel electrophoresis due to the generation of a 

sufficient cell mass.

IIIC.4.2 Molecular Characterization

Within the last decade, a variety of culture-independent 

genetic analyses have been used to complement traditional 

culture-dependent methods (enrichment and isolation). 

Many of these molecular biological methods rely on 16S 

rDNA sequences, including in-situ hybridization (Amann 

et al., 1990; Amann et al., 1995), direct amplification of 

16S rDNA, and additional analysis using community “DNA 

fingerprinting” such as temperature gradient gel electrophoresis (TGGE) (Felske et al., 1998), denaturing gradient gel 

electrophoresis (DGGE) (Muyzer et al. 1993), restriction 

fragment length polymorphism (RFLP) (Martinez-Murcis 

et al., 1995), single-strand conformation polymorphism 

(SSCP) (Lee et al., 1996), terminal RFLP (Clement et al., 

1998), or 16S rDNA cloning-sequencing (Wise et al., 1997). 

To date, most of the results obtained by using molecular 
71

techniques have been provided by cloning-sequencing of 

16S rDNA fragments. Although 16S rDNA cloning-sequencing is successful on the reconnaissance of microbial 

diversity by detecting infrequent sequences from various 

habitats, thereby avoiding limitations of traditional cultivation techniques, it is time-consuming and problematic for 

multiple sample analysis. Since the cloning approach 

cannot provide an immediate overview of the community 

structure, many environmental laboratories apply DGGE 

to detect population shifts. DGGE can be used for simultaneous analysis of multiple samples obtained at various 

time intervals to detect microbial community changes; an 

advantageous feature in studying microbial ecology and 

MNA.

These methods characterize differing aspects of the subsurface microbial community, which may be used alone or 

in combination to further delineate the impact of microbial 

processes on ground-water chemistry. DGGE provides a 

simple approach to obtaining profiles of microbial communities and identifying temporal and spatial variations which 

occur in response to various environmental conditions 

(Muyzer et al., 1993). It is also possible to infer the phylogeny of community members by DNA sequence analysis of 

re-amplified fragments, after they are excised from the gel, 

where bands corresponding to each microorganism can be 

separated through DGGE. Fluorescent in-situ hybridization (FISH) provides a powerful tool for directly studying 

organisms within the environment by providing information 

on cell morphology, phylogenic affiliation and the ability to 

quantify organisms (Amann et al.,1990; Amann et al., 1995). 

During the last few years numerous efforts have been made 

to increase the sensitivity of FISH, including multi-labeled 

polynucleotide probes (Pernthaler et al., 2002). The RFLP 

approach involves electrophoretic analysis where DNA is 

detected with probes after Southern blotting. RFLP is applied broadly since it has a good predictive power and can 

rapidly identify phylogenic relatedness of clusters of very 

closely related or even identical strains. It can screen large 

numbers of isolates to identify a much smaller subset of 

representative types to be resolved by 16S rRNA sequencing. Although RFLP generates a complex set of rDNA 

bands that can be used to group closely related strains, it 

does not provide the distance between strains that are not 

closely related, as does rRNA sequencing. In contrast, the 

T-RFLP approach uses restriction enzymes, coupled with 

PCR, in which only fragments containing a fluorescent 

tag are detected. The use of T-RFLP is advantageous as 

a rapid screening tool and does not require culturing or a 

genetic database.

IIIC.4.3 Sampling Considerations

Hydrogeologic conditions are of overriding significance 

in designing and conducting subsurface microbiological 

sampling programs because the selection of equipment and 

the location of sampling points are necessarily subject to 

site-specific conditions. With regard to the type of sample 

(core or water), core samples provide more information in 

defining the horizontal and vertical distribution of microbes 

even though they are intrinsically disruptive and prohibit 

repeated sampling at the same location. Ground-water 

samples, on the other hand, can be obtained from the same 

well repeatedly but may not quantitatively or qualitatively 

reflect conditions in the aquifer. Since there are uncertainties that mandate caution in the extrapolation of information 

obtained from either type of sample, a thorough characterization may require both water and core samples. 

In studies addressing the origin and nature of subsurface 

microbes, an important consideration is the extent organisms cultured or manipulated in the laboratory represent 

the intrinsic microbial community. Therefore, microbiologists 

are challenged when collecting not only representative but 

also microbially uncontaminated samples. To minimize 

contamination, after a core is obtained using strict aseptic 

methods, care should be taken not to disturb or contaminate 

the sample. Processing should be performed as quickly 

as possible under anaerobic and aseptic conditions while 

in the field. Surface layers of the core should be scraped 

away using a sterile sampling devise and discarded so that 

only the center of the core is packaged in doubled sterile 

sample bags. The portion used for DNA/PLFA analyses 

must be rapidly frozen with liquid nitrogen and stored at 

–70 L°C. Another portion should be flushed with inert 

gases (N2

 or Ar), sealed in canning jars and placed inside 

cans containing oxygen-scavenging catalyst packets (Gas 

Pak; BBL Inc., Franklin Lakes, N.J.), and stored at 4 °C for 

microbial counts and cultural techniques for analysis within 

24 hours. Frozen (-70 °C) genomic DNA can be extracted 

from core sample using an UltraClean soil DNA kit (MoBio. 

Solana, CA). 

Ground-water samples will be unfiltered for microbial counts 

and cultural techniques and filtered for DNA/PLFA analysis. 

Community analysis involving “DNA fingerprinting” requires 

ultrafiltration of ground water (50 to 100 L) using inorganic 

(AnodiscTM, Whatman) filters (0.2 μm pore size). The filter 

is placed in sterile bag and rapidly freeze dried with liquid 

nitrogen. Since PCR bias can be introduced, particularly 

during cell lysis and PCR amplification, a physical method 

such as bead mill homogenization should be used to effectively lyse all cell types, including those that are most 

recalcitrant to physical and enzymatic treatments (More 

et al., 1994).

IIIC.5 Implications for Natural Attenuation 

Assessment

Ultimately, the extent and degree of site characterization 

to define the redox status of the aquifer will represent a 

balance between technical information needs and the 

cost associated with the different proposed data collection or evaluation schemes. The primary objective of site 

characterization is to identify the mechanism leading to 

contaminant attenuation at a given site. Emphasis of this 

characterization effort should be given to direct measurements of ground water conditions and solid phase characteristics of the aquifer that result from these conditions. 

Measurements and/or tests conducted with subsurface 

samples retrieved within the zones where attenuation occurs will provide the most direct means to evaluate on-going 

reaction processes. This knowledge will guide approaches 
72

to assess the capacity of the aquifer to sustain contaminant 

attenuation within the plume and to evaluate the longterm stability of immobilized contaminants. Evaluations 

conducted on subsurface samples also have the benefit 

of implicitly incorporating characteristics/factors of ground 

water and aquifer solids that may be difficult to adequately 

parameterize within analytical models.

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