Monitored Natural Attenuation 

of Inorganic Contaminants in 

Ground Water

Volume 2 

Assessment for Non-Radionuclides 

Including Arsenic, Cadmium, Chromium, 

Copper, Lead, Nickel, Nitrate, 

Perchlorate, and Selenium

EPA/600/R-07/140

October 2007

Monitored Natural Attenuation of Inorganic 

Contaminants in Ground Water

Volume 2 

Assessment for Non-Radionuclides Including 

Arsenic, Cadmium, Chromium, Copper, Lead, 

Nickel, Nitrate, Perchlorate, and Selenium

Edited by

Robert G. Ford, Richard T. Wilkin, & Robert W. Puls

U.S. Environmental Protection Agency 

Office of Research and Development 

National Risk Management Research Laboratory

Ada, Oklahoma 74820

Project Officer

Robert G. Ford

Ground Water and Ecosystems Restoration Division

National Risk Management Research Laboratory

Ada, Oklahoma 74820

National Risk Management Research Laboratory

Office of Research and Development

U.S. Environmental Protection Agency

Cincinnati, Ohio 45268
ii

Notice

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

Development managed portions of the technical work described here under EPA 

Contract No. 68-C-02-092 to Dynamac Corporation, Ada, Oklahoma 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

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

Figures ......................................................................................................................................... vii

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

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

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

Chapter 1 - Cadmium .....................................................................................................................1

Occurrence and Distribution ....................................................................................................1

Geochemistry and Attenuation Processes ..............................................................................1

Site Characterization ................................................................................................................4

Long-term Stability and Capacity .............................................................................................5

Tiered Analysis ........................................................................................................................6

References ...............................................................................................................................7

Chapter 2 - Lead ..........................................................................................................................11

Occurrence and Distribution ..................................................................................................11

Geochemistry and Attenuation Processes ............................................................................12

Site Characterization ..............................................................................................................15

Long-term Stability and Capacity ...........................................................................................16

Tiered Analysis ......................................................................................................................17

References .............................................................................................................................18

Chapter 3 - Nickel .........................................................................................................................21

Occurrence and Distribution ..................................................................................................21

Geochemistry and Attenuation Processes ............................................................................22

Site Characterization ..............................................................................................................25

Long-term Stability and Capacity ...........................................................................................27

Tiered Analysis ......................................................................................................................27

References .............................................................................................................................28

Chapter 4 - Copper ......................................................................................................................33

Occurrence and Distribution ..................................................................................................33

Geochemistry and Attenuation Processes ............................................................................34

Site Characterization ..............................................................................................................36

Long-term Stability and Capacity ...........................................................................................37

Tiered Analysis ......................................................................................................................37

References .............................................................................................................................38

Chapter 5 - Chromium ..................................................................................................................43

Occurrence and Distribution ..................................................................................................43

Geochemistry and Attenuation Processes ............................................................................43

Site Characterization ..............................................................................................................47

Long-term Stability and Capacity ...........................................................................................50

Tiered Analysis ......................................................................................................................50

References .............................................................................................................................51

Chapter 6 - Arsenic ......................................................................................................................57

Occurrence and Distribution ..................................................................................................57

Geochemistry and Attenuation Processes ............................................................................58

Site Characterization...............................................................................................................61
vi

Long-term Stability and Capacity ...........................................................................................64

Tiered Analysis ......................................................................................................................64

References .............................................................................................................................66

Chapter 7 - Selenium ...................................................................................................................71

Occurrence and Distribution ..................................................................................................71

Geochemistry and Attenuation Processes ............................................................................73

Site Characterization ..............................................................................................................75

Long-term Stability and Capacity ...........................................................................................79

Tiered Analysis ......................................................................................................................79

References .............................................................................................................................80

Chapter 8 - Nitrate ........................................................................................................................87

Occurrence and Distribution ..................................................................................................87

Geochemistry and Attenuation Processes ............................................................................91

Site Characterization ..............................................................................................................92

Long-term Capacity ...............................................................................................................94

Tiered Analysis ......................................................................................................................95

References .............................................................................................................................97

Chapter 9 – Perchlorate .............................................................................................................101

Occurrence and Distribution ................................................................................................101

Geochemistry and Attenuation Processes ..........................................................................102

Site Characterization ............................................................................................................103

Long-term Capacity .............................................................................................................105

Tiered Analysis ....................................................................................................................105

References ...........................................................................................................................106
vii

Figures

Figure 1.1 Species distribution of Cd(II) in pure water as a function of pH at 25 °C. ...................2

Figure 1.2 Cadmium speciation as a function of pH in solution containing chloride (100 mg L-1), 

 sulfate (100 mg L-1), and inorganic carbon (100 mg L-1). ............................................2

Figure 1.3 Eh-pH diagram for cadmium (total Cd = 10-5 molal, total C = 10-3 molal, total S = 10-3 

molal; all organic cadmium complexes are suppressed; activity coefficients for all 

species are set equal to 1). .........................................................................................3

Figure 2.1 Pb(II) species distribution in pure water at 25° C. .....................................................12

Figure 2.2 Species distribution of lead in solution with 100 mg L-1 chloride, 100 mg L-1 sulfate, 

and 100 mg L-1 total inorganic carbon, based on thermodynamic data in 

MINTEQA2 (Allison et al., 1990). ..............................................................................12

Figure 2.3. Eh-pH diagram for lead (total Pb = 10-5molal, total C = 10-3molal, total S = 10-3molal; 

all organic lead complexes are suppressed; activity coefficients for all species are 

set equal to 1). ...........................................................................................................13

Figure 2.4 Pb(II) activity in equilibrium with PbCO3

 (at total inorganic carbon equal to 

0.001 molal), PbSO4 (at total sulfate equal to 0.1 molal), and PbS (at total sulfide 

equal to 0.001 molal). ................................................................................................13

Figure 3.1 (a) Predicted solubility of various Ni precipitates that could form in aerobic ground 

water with concentrations of Al and Si controlled by the solubility of the clay 

mineral, kaolinite. .......................................................................................................23

Figure 3.2 Eh-pH diagrams for nickel at 25 °C. (a) System Ni-H2

O-Ca-Al-NO3-HCO3-SO4 

(2 mg Ni/L; 40 mg Ca/L; 3 mg Al/L; 6 mg NO3/L; 60 mg HCO3/L; 100 mg SO4/L). ....23

Figure 3.3 Nickel sorption as a function of pH in the presence of an hypothetical aquifer 

sediment with iron and manganese oxides reflective of the crustal abundance of 

these elements. ..........................................................................................................24

Figure 4.1 Solubility of copper oxide and copper hydroxide as a function of pH in the system 

Cu-O-H at 25 °C. .......................................................................................................34

Figure 4.2 Eh-pH diagram for copper at 25 °C (total inorganic carbon = 10-2 molal; total 

sulfur = 10-3 molal; total copper = 10-5 molal). ............................................................35

Figure 4.3 Solubility and speciation of copper as a function of pH and log fugacity of CO2

(gas) 

at 25 °C. .....................................................................................................................35

Figure 5.1 Distribution of 1 mM (millimoles per liter, which equals approximately 52 mg L-1) 

chromium(VI) plotted as a function of pH in equilibrium with barite. .........................44
viii

Figure 5.2 Calculated distribution of 1 mM Cr(VI) (approximately 52 mg L-1) in the presence 

of 1 mM K (approximately 39 mg L-1) and hydrous ferric oxide calculated as a 

function of pH. ............................................................................................................45

Figure 5.3 Concentration of dissolved Cr(VI) in equilibrium with Cr(VI) adsorbed on freshly 

precipitated hydrous ferric oxide. ...............................................................................45

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

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

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

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

systems. .....................................................................................................................60

Figure 7.1 Eh-pH diagram for selenium at 25 °C using thermodynamic data from 

Séby et al. (2001). ......................................................................................................73

Figure 7.2 Eh-pH diagram for selenium at 25 °C using thermodynamic data from 

Séby et al. (2001). ......................................................................................................74
ix

Tables

Table 1.1 Natural attenuation and mobilization pathways for cadmium. .....................................4

Table 2.1 Natural attenuation and mobilization pathways for lead. ...........................................15

Table 3.1 Natural attenuation and mobilization pathways for nickel. .........................................25

Table 4.1 Natural attenuation and mobilization pathways for copper. .......................................37

Table 5.1 Natural attenuation and mobilization pathways for chromium. ..................................47

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

aqueous samples. ......................................................................................................48

Table 6.1 Natural attenuation and mobilization pathways for arsenic. ......................................62

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

Table 7.1 Natural attenuation and mobilization pathways for selenium. ....................................76

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

Table 7.3 Review of selenium isotope fractionation ranges for abiotic-biotic processes 

during reduction and oxidation (Johnson, 2004). ......................................................78

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

Table 8.2 Natural attenuation and mobilization pathways for nitrate. ........................................90

Table 9.1 Natural attenuation and mobilization pathways for perchlorate. ..............................103

Table 9.2 Published USEPA methods for determination of perchlorate and other 

Cl-bearing inorganic ions in aqueous samples. .......................................................104
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 contaminant chapters to which contributions were made:

Contributing 

Author Chapter Affiliation

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

Lead

Nickel

Copper

Arsenic

Selenium

Perchlorate

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

45268

Arsenic

Selenium

Nitrate

Perchlorate

Chunming Su Selenium USEPA/ORD, National Risk Management Research Laboratory, Ada, OK 74820

Nitrate

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

Kirk G. Scheckel Nickel USEPA/ORD, National Risk Management Research Laboratory, Cincinnati, OH 

45268

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

Chromium

Arsenic

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

Albuquerque, New Mexico 87185

Perchlorate

Critical and constructive reviews were provided by Kenneth Lovelace (USEPA/OSRTI), John Washington (USEPA/

ORD National Exposure Research Laboratory, Athens, GA), Jackson Ellington (USEPA/ORD National Exposure 

Research Laboratory, Athens, GA), George Redden (Idaho National Laboratory), and Sue Clark (Washington State 

University). Pat Bush (Ada, OK) is acknowledged for her technical editing to provide consistency in formatting and 

grammar. Martha Williams (Contract #68-W-01-032) assisted with final editing and formatting for publication.
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 second volume of a set of three volumes that address the technical basis and 

requirements for assessing the potential applicability of MNA as part of a ground-water remedy for plumes with nonradionuclide and/or radionuclide inorganic contaminants. Volume 2, titled “Assessment for Non-Radionuclides Including Arsenic, Cadmium, Chromium, Copper, Lead, Nickel, Nitrate, Perchlorate, and Selenium,” consists of individual 

chapters that describe 1) the natural processes that may result in the attenuation of the listed contaminants and 2) 

data requirements to be met during site characterization. Emphasis is placed on characterization of immobilization 

and/or degradation processes that may control contaminant attenuation, as well as technical approaches to assess 

performance characteristics of the MNA remedy. A tiered analysis approach is presented to assist in organizing 

site characterization tasks in a manner designed to reduce uncertainty in remedy selection while distributing costs 

to address four primary issues:

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

2. Determination of the rate and mechanism of attenuation;

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

and 

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

establishing a contingency plan.

Where feasible, Agency-approved analytical protocols currently implemented for waste site characterization are 

identified, along with modifications that may be warranted to help insure the quality of site-specific data. In situations where Agency methods or protocols are unavailable, recommendations are made based on review of the 

existing technical literature. It is anticipated that future updates to these recommendations may be warranted with 

increased experience in the successful application of MNA as part of a ground-water remedy and the development 

of new analytical protocols. 

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

monitoring system design in fractured rock, karst, and other such highly heterogeneous settings is beyond the 

scope of this document. Ground water and contaminants often move preferentially through discrete pathways (e.g., 

solution channels, fractures, and joints) in these settings. Existing techniques may be incapable of fully delineating the pathways along which contaminated ground water migrates. This greatly increases the uncertainty and 

costs of assessments of contaminant migration and fate and is another area of continuing research. As noted in 

OSWER Directive 9200.4-17P, “MNA will not generally be appropriate where site complexities preclude adequate 

monitoring.” The directive provides additional discussion regarding the types of sites where the use of MNA may 

be appropriate.

This document focuses on monitoring the saturated zone, but site characterization and monitoring for MNA or any 

other remedy typically would include monitoring of all significant pathways by which contaminants may move from 

source areas and contaminant plumes to impact receptors (e.g., surface water and indoor air). 
xii

Nothing in this document changes Agency policy regarding remedial selection criteria, remedial expectations, or the 

selection and implementation of MNA. This document does not supercede any guidance. It is intended for use as a 

technical reference in conjunction with other documents, including OSWER Directive 9200.4-17P, “Use of Monitored 

Natural Attenuation at Superfund, RCRA Corrective Action, and Underground Storage Tank Sites” (http://www.epa.

gov/swerust1/directiv/d9200417.pdf).
1

Cadmium

Richard T. Wilkin

Occurrence and Distribution

Cadmium is comparatively rare in the environment with an 

average abundance similar to other second- and third-row 

transition metals (e.g., silver and mercury). The median 

concentration of cadmium in soils and sediments ranges 

from about 0.04 to 1.8 mg kg-1 (Reimann and Caritat, 

1998). Where cadmium concentrations are elevated, it is 

typically found in association with sulfide ores of zinc and 

sometimes with ores of copper and lead. The primary 

mineral associations of cadmium are with otavite (CdCO3), 

greenockite (CdS), sphalerite (ZnS), smithsonite (ZnCO3), 

and hemimorphite (Zn4Si2O7(OH)2·H2O). Soil weathering 

can lead to release of the Cd2+ ion, which is generally 

soluble and mobile in water.

The primary industrial uses of cadmium are metal plating, 

production of Ni-Cd batteries, as a stabilizer in plastics, and 

as a pigment. According to Minerals Information statistics 

for 2001, approximately 75% of the U.S. apparent consumption of cadmium (a total of about 2.4 million pounds) 

went into production of Ni-Cd batteries (Wolke, 2003). 

The largest sources of cadmium contamination to ground 

water and surface water are from sewage sludge, mines 

(e.g., mine water, mine tailings leachate), metal smelters 

(process waters), battery recycling plants, and wastes from 

electroplating facilities.

Remedial Technologies

Treatment of cadmium and other heavy metals in industrial 

wastewater streams is often achieved by precipitation using 

lime, sodium carbonate, alkaline sulfides, or organosulfides. These treatment methods are generally unsuitable 

for drinking water. Ion exchange resins and adsorption 

substrates are in most cases used for treatment of drinking water contaminated with cadmium (Zhao et al., 2002; 

Lai et al., 2002). Technology classes suitable for remediation of cadmium-contaminated soils include containment, 

solidification/stabilization, and separation/concentration 

(USEPA, 1997). Containment technologies applied at metal 

contamination sites to minimize the transport of cadmium 

and co-contaminants out of source zones include caps 

and vertical barriers. Reactive barriers are appropriate for 

treatment of some cadmium and co-contaminant ground 

water plumes (e.g., Gibert et al., 2003; Wang and Reardon, 

2001).

Regulatory Aspects

Cadmium and its compounds are very toxic to nearly all living organisms. The EPA has set the maximum contaminant 

level (MCL) for cadmium in drinking water at 0.005 mg L-1. 

Cadmium is fairly mobile and soluble in water at low to 

near-neutral pH. The main routes by which cadmium 

enters the human body are ingestion of plant-based food 

and inhalation of cadmium-bearing dusts. The kidney is 

the primary organ affected by exposure to cadmium. For 

non-potable water sources, ambient water quality criteria 

(AWQC) that are protective of aquatic life may serve as 

alternative cleanup goals. For cadmium, current statutes 

list both acute and chronic criteria for fresh waters as 

0.002 mg L-1 and 0.00025 mg L-1, respectively, for a water 

hardness of 100 mg L-1 (USEPA, 2006; http://www.epa.

gov/waterscience/criteria/nrwqc-2006.pdf). Adjustments 

to these criteria are to be applied for waters with different 

hardness. An example of where this criterion may apply 

is a site where contaminated ground water discharges to 

surface water.

Geochemistry and Attenuation Processes

Aqueous Speciation

Dissolved forms of cadmium are only present in the +2 

valence state. Cadmium has a tendency to form aqueous complexes with both inorganic and organic ligands, 

although the uncomplexed Cd2+ ion is fairly stable. The 

most important inorganic cadmium complexes are with hydroxide, chloride, sulfate, bicarbonate, carbonate, cyanide, 

and ammonia. Complexation of cadmium with humic acids 

is important under conditions of high dissolved organic 

carbon (DOC) concentrations, but binding of cadmium with 

humic acids appears to be weaker when compared to lead 

(Abate and Masini, 2002; Christensen and Christensen, 

1999; Dunnivant et al., 1992). In highly reducing systems, 

cadmium complexation with bisulfide is possible. It is likely 

that cadmium toxicity is related to its strong tendency to 

form bonds with thiol functional groups in certain enzymes 

which results in the displacement of biologically essential 

metals (Baes and Mesmer, 1976).

The fractional abundance of Cd-OH species in water as a 

function of pH is shown in Figure 1.1. The distribution diagram for cadmium hydroxy complexes indicates that Cd2+, 

Cd(OH)+, Cd(OH)2

0 are the most significant species below 
2

pH 12. At low total concentrations of cadmium, hydrolysis of 

Cd2+ becomes significant above about pH 9. Complexation 

of cadmium with chloride and ammonia becomes important 

as aqueous concentrations of these ligands exceed 10-2

molal and 10-3 molal, respectively (e.g., Lindsay, 1979).

Figure 1.1 Species distribution of Cd(II) in pure water 

as a function of pH at 25 °C.

Figure 1.2 shows the fractional abundance of cadmium 

species as a function of pH in an aqueous solution containing a mixture of chloride, sulfate, and inorganic carbon. Again the uncomplexed Cd2+ ion dominates below 

pH 8.5. In general, complexation of cadmium with chloride 

and sulfate is most important at near-neutral to low pH; 

carbonate complexation is most important at pH 9 to 11; 

and, hydroxyl species dominate cadmium complexation at 

pH > 11 (Figure 1.2).

The identity of cadmium bisulfide complexes and their 

formation constants have been discussed by Daskalakis 

and Helz (1992) and Wang and Tessier (1999). Uncertainty persists regarding the stoichiometry of the most 

important cadmium complexes in sulfidic waters. This 

uncertainty mainly stems from the experimental approach 

that has been traditionally used to extract thermodynamic 

data, i.e., evaluation of CdS solubility over a range of total 

cadmium concentrations, total sulfide concentrations, and 

pH. Data presented in Wang and Tessier (1999) indicate 

that Cd(HS)2

0 is the dominant species at ΣS(-II)=10-5 molal 

and over the pH range from 6 to 8, typically encountered in 

natural sulfidic waters. At lower total sulfide concentrations, 

CdHS+ and Cd2+ become increasingly important (Wang and 

Tessier, 1999). 

Figure 1.2 Cadmium speciation as a function of pH in 

solution containing chloride (100 mg L-1), 

sulfate (100 mg L-1), and inorganic carbon 

(100 mg L-1). Cadmium chloride complexes 

include CdCl+ and CdCl2

0. Cadmium sulfate 

complexes include CdSO4

0 and Cd(SO4)22-. 

Cadmium complexes with inorganic carbon 

include CdCO3

0, CdHCO3+, and Cd(CO3)22-

. 

Cadmium hydroxy complexes include 

CdOH+, Cd(OH)2

0, and Cd(OH)3-

. Total 

cadmium is equal to 1 mg L-1.

Solubility

An Eh-pH diagram for cadmium is shown in Figure 1.3. 

Inspection of this diagram indicates that at the specified 

conditions Cd2+ is the soluble form of cadmium at pH < 5 

and at moderate to highly oxidizing redox potentials. At 

near-neutral to moderately alkaline pH (6 to about 12) 

cadmium carbonate (otavite) is stable, and at pH > 12.5 

cadmium hydroxide is stable. In sulfidic environments, 

cadmium sulfide (greenockite) is stable over a wide pH 

range. Solubility expressions for cadmium carbonate and 

cadmium sulfide are given by:

CdCO3 + H+ = Cd2+ + HCO3

- (log K= -0.9)

and

CdS + H+ = Cd2+ + HS- (log K = -14.4)

In natural deposits, cadmium often substitutes for zinc in 

the mineral structures of sphalerite (ZnS) and smithsonite 

(ZnCO3). O’Day et al. (1998) suggest that as cadmiumsubstituted sphalerite weathers, cadmium is preferentially 

partitioned into the aqueous phase over zinc. Zinc was 

found to form various zinc hydroxides and/or zinc-iron 

oxyhydroxides depending on the total amount of iron in 

the system. Cadmium was not identified in the solid-phase 

products from weathering indicating its general tendency 

to be mobile in the aqueous phase (O’Day et al., 1998; 

Carroll et al., 1998).
3

Figure 1.3 Eh-pH diagram for cadmium (total Cd = 

10-5 molal, total C = 10-3 molal, total S = 10-3 

molal; all organic cadmium complexes are 

suppressed; activity coefficients for all species are set equal to 1).

Cadmium carbonates - Cadmium is known to form solid-solutions with calcium carbonate (calcite). The Cd2+ and Ca2+

ions are nearly the same size with crystal radii of 1.09 and 

1.14 angstroms, respectively, assuming octahedral coordination. The formation of Ca-Cd carbonate solid solutions 

is environmentally significant because solid solutions are 

generally more stable and less soluble than endmember 

compositions. Cadmium uptake from aqueous solution by 

calcite has been widely studied (e.g., McBride, 1980; Davis 

et al., 1987; Papadopoulos and Rowell, 1988; Zachara et 

al., 1991; Stipp et al., 1992, 1993; Tesoriero and Pankow, 

1996; Chiarello et al., 1997; Martin-Garin et al., 2003). 

Cadmium uptake is thought to consist of two processes. 

The first process is rapid sorption and the second process 

is incorporation into the crystal lattice and formation of an 

otavite-calcite solid-solution. The latter process has been 

confirmed through determination of cadmium solid phase 

speciation during reaction with calcite (Bailey et al., 2005). 

Cadmium partitioning to carbonaceous materials has been 

applied by Wang and Reardon (2001) for the removal of 

cadmium from wastewater streams.

Cadmium phosphates - Santillan-Medrano and Jurinak 

(1975) observed the formation of cadmium phosphate 

precipitates in soil systems containing phosphorus. Precipitation of phosphate compounds over carbonate compounds was favored in phosphorous-containing systems at 

pH < 7.5. Substitution of cadmium into natural apatite has 

also been documented (Sery et al., 1996) and may be a 

more common route for partitioning to phosphate minerals 

at concentrations undersaturated with respect to precipitation of cadmium phosphate.

Cadmium hydroxides – Baes and Mesmer (1976) report the 

existence of three crystalline forms of Cd(OH)2. At 25° C 

the stable form is β-Cd(OH)2. This material is fairly soluble 

at circumneutral pH. Cadmium concentrations below the 

MCL of 0.005 mg L-1 would only be approached at pH > 11 

in systems saturated with respect to β-Cd(OH)2.

Cadmium sulfides – It is well known that Cd2+ and HS- react 

to form a very insoluble, yellow-colored precipitate. However, Daskalakis and Helz (1992) point out that under some 

conditions dissolved cadmium bisulfide complexes are very 

stable; consequently, the effectiveness of sulfide treatment 

for cadmium in wastewater has been overestimated in some 

cases due to uncertainty regarding the stability of cadmium 

bisulfide species.

Framson and Leckie (1978) examined the limit of coprecipitation of cadmium and ferrous monosulfide. Their experimental data suggest only limited coprecipitation, likely due 

to the size mismatch between the ferrous (~0.75 angstroms) 

and cadmium (~1.09 angstroms) cations. They suggest that 

in sulfidic systems, cadmium precipitates primarily through 

surface exchange with ferrous monosulfide substrates or 

as unsubstituted cadmium sulfide. Parkman et al. (1999) 

performed X-ray absorption spectroscopy experiments 

and also concluded that a CdS phase is formed as cadmium interacts with iron monosulfide. On the other hand, 

Coles et al. (2000) found up to 29% replacement of iron 

by cadmium in mackinawite (FeS). It is possible that this 

high percentage of replacement occurs at the surface of 

very fine-grained iron sulfide particles. Coles et al. (2000) 

found that the mixed ferrous-cadmium sulfides are more 

insoluble than pure mackinawite.

Adsorption

Adsorption/desorption behavior of cadmium is strongly 

a function of pH, and to a lesser extent a function of the 

solution concentration of cadmium and the concentration 

of competing cations or complexing ligands. At low concentrations of cadmium, sharp adsorption edges provide 

evidence that cadmium forms strong bonds with mineral 

surfaces. In general, the presence of calcium and magnesium reduces the extent of cadmium removal by aquifer 

solids. The presence or addition of zinc, which tends to be 

more strongly adsorbed, can reduce the amount of cadmium 

uptake by iron and aluminum oxides, indicating that zinc 

competes for similar adsorption sites and is preferentially 

adsorbed over cadmium (Benjamin and Leckie, 1980). It 

has also been observed that cadmium may preferentially 

adsorb to manganese oxides when they are present in 

sufficient quantities (Bellanca et al., 1996). Tonkin et al. 

(2004) have evaluated published cadmium adsorption data 

to determine surface complexation constants that may be 

employed to assess the potential extent of adsorption onto 

manganese oxides.

Ainsworth et al. (1994) examined the sorption behavior of 

cadmium on freshly prepared and aged hydrous ferric oxide 

(HFO). Their results indicate that HFO effectively removes 

cadmium from solution at pH above about 6.7. In general, 

the pH-dependent adsorption behavior parallels the change 
4

in aqueous speciation from Cd2+ to cadmium hydroxide 

species (see Fig. 1.1), although adsorption occurs at pH 

values where cadmium hydroxide species are unexpected 

in bulk solution. Aging times of up to 21 weeks showed little 

effect on the sorption behavior of cadmium onto HFO. HFO 

aged in the presence of Cd2+ ions showed some desorption 

hysteresis suggesting that cadmium is incorporated into the 

metal oxide structure during recrystallization. Martínez and 

McBride (1998) suggest that coprecipitation of cadmium 

with amorphous iron oxides results in more reduced concentrations than can be achieved through surface adsorption alone. However, Ford et al. (1997) report that during 

long-term aging of hydrous iron oxides, cadmium desorbs 

or is released suggesting minimal incorporation of cadmium 

into the goethite or hematite structures.

Lai et al. (2002) investigated the adsorption characteristics 

of cadmium and humic acid onto iron oxyhydroxide-coated 

quartz sands. The adsorption of both cadmium and humic 

acid was highly pH dependent. Cadmium adsorption increased with pH, whereas humic acid adsorption decreased 

as pH increased. The presence of humic acid was found 

to result in increasing cadmium adsorption capacity in the 

pH range of 4-6.

Redox Chemistry

In natural systems cadmium is present in the +2 oxidation 

state. Therefore, the geochemical transport processes of 

cadmium are not directly tied to changes in redox conditions. Because cadmium forms stable precipitates and 

aqueous complexes with redox-sensitive elements such as 

sulfur and carbon, its mobility potential is indirectly tied to 

redox conditions. In sulfate-reducing systems, cadmium is 

expected to form insoluble CdS precipitates or coprecipitates with FeS (DiToro et al., 1990). In moderately reducing 

but non-sulfidic systems, however, reductive dissolution of 

hydrous ferric oxides with adsorbed cadmium could result 

in cadmium mobilization.

Several studies indicate that concentrations of dissolved 

cadmium increase when reduced systems are oxidized, 

such as when dredged sediments are land filled. This behavior may be due to oxidative dissolution of metal sulfides 

or due to the decomposition of organic materials that bind 

cadmium (e.g., Cooper and Morse, 1998; Simpson et al., 

2000; Martínez et al., 2002).

Colloidal Transport

Transport of cadmium via colloids can be significant in 

ground water and surface water systems. Both mineral and 

organic particles can play a role in binding and transporting 

cadmium. Cadmium adsorbed to colloidal hydrous ferric 

oxides may subsequently desorb due to pH decreases or 

due to decreases in the oxidation-reduction potential.

Site Characterization

Cadmium mobility in ground water is governed by the total 

concentration of cadmium, the distribution of cadmium species in water, and the nature of cadmium partitioning in the 

solid phase. The development of site-conceptual models 

for predicting the long-term fate of cadmium at a contaminated site will require information on the distribution and 

concentration of cadmium in the aqueous phase and the 

solid phase. Table 1.1 indicates possible natural attenuation and mobilization pathways for cadmium. Details of the 

types of analytical measurements that may be conducted 

on sampled ground water and aquifer sediments to assist 

in identifying the attenuation mechanism(s) are discussed 

in the following paragraphs.

Table 1.1 Natural Attenuation and Mobilization Pathways for Cadmium

Attenuation Processes Mobilization Processes Characterization Approach

Precipitation of insoluble 

carbonates, sulfides, and 

hydroxides. In general, pH > 

8 will drive precipitation reactions resulting in Cd concentrations below the MCL.

Dissolution of carbonates at low pH; 

oxidative dissolution of sulfides at low to 

neutral pH and high Eh; degradation of 

organic matter; complexation/stabilization in the presence of DOC.

Evaluation of cadmium concentration in the 

aqueous phase. Determination of total Cd in the 

solid matrix. Evaluation of solid phase partitioning using sequential extraction methodologies. 

Evaluation of long-term storage capacity.

Sorption (reversible) to iron 

hydroxides, organic matter, 

carbonates, sulfides (pH > 6 

to 8). Substitution for Zn in 

minerals.

Desorption at low pH; complexation/stabilization in the presence of DOC. Reductive dissolution of iron hydroxides. 

Oxidation of metal sulfides.

Evaluation of cadmium speciation in the aqueous phase. Determination of total Cd in the solid 

matrix. Evaluation of solid phase partitioning 

using sequential extraction methodologies. 

Batch and column testing to determine Cd uptake capacity of site-specific aquifer materials 

at variable geochemical conditions.
5

Aqueous Measurements

Quantitative measurement of cadmium concentrations in 

aqueous solutions is typically carried out using inductively 

coupled plasma atomic emission spectroscopy (ICP-AES), 

inductively coupled plasma mass spectroscopy (ICP-MS), 

or atomic absorption spectroscopy (AAS). Input data to 

geochemical codes (e.g., MINTEQA2, PHREEQC, EQ3/6) 

for determining aqueous speciation also require, at a 

minimum, concentrations of major anions, major cations, 

dissolved organic carbon, temperature, and pH. In addition, while regulatory requirements stipulate that unfiltered 

ground-water samples be analyzed to support regulatory 

decisions at a contaminated site, it may be necessary to 

also collect filtered samples to help define the process(es) 

controlling contaminant mobility. The use of 0.45 μm pore 

size filters is common as an arbitrary cutoff point to differentiate between dissolved and particulate phases in 

water samples. However, caution is recommended when 

using this approach, particularly for Fe and Al and other 

elements that may be associated with Fe or Al particles 

(including Cd) that could pass through 0.45 μm filters. The 

use of filters with pore sizes less than 0.1 μm will generally 

provide a better assessment of the dissolved vs. particulate 

load in ground water.

Solid Phase Measurements

The implementation of an analytical approach to identify 

cadmium speciation in aquifer sediments is a challenging 

process. The accuracy of the analytical finding is dependent on the method of sample collection/preservation 

and the tools used to identify the mechanism of cadmium 

partitioning. It is recommended that the analytical protocol 

be designed to address the potential redox sensitivity of 

the solid phase(s) to which cadmium may be partitioned 

(e.g., sulfides in reduced sediments). Tools to evaluate the 

mechanism of cadmium solid phase partitioning range in 

complexity from relatively simple chemical extractions to 

advanced spectroscopic techniques.

The total concentration of cadmium in soils, sediments, and 

aquifer materials may be determined by X-ray fluorescence 

spectroscopy (XRF), or by ICP-AES after digestion in 

mineral acids. A variety of digestion or extraction methods 

can be found in the literature (Amacher, 1996). Neutron 

activation analysis is not commonly employed due to the 

scarcity of neutron sources required to irradiate the sample. 

X-ray fluorescence is the most attractive approach due 

to the relative ease of sample preparation, which may be 

conducted with the sample in its original state or following 

fusion with lithium metaborate. When combined with the 

determination of other major or trace elements in the solid 

sample, this provides an initial step for assessing possible 

association of cadmium with various solid phase components. This type of analysis can be conducted on the bulk 

sample as well as at a microscopic level using wavelength 

(electron microprobe) or energy dispersive spectroscopy 

coupled to a scanning or transmission electron microscope. 

Microscopic examination allows one to better differentiate 

whether cadmium may be distributed across a number of 

different mineral phases within the solid sample or primarily associated with a discrete phase. There are limitations 

to this approach (Pye, 2004), a significant one being that 

the analysis does not necessarily provide unique mineral 

identification necessitating the collection of supporting 

mineralogical and chemical data.

More detailed information on the specific partitioning 

mechanism(s) controlling cadmium solid phase speciation 

is typically required to adequately support site assessment for potential reliance on natural attenuation as part 

of a site remedy. There have been many applications of 

sequential extraction schemes to assess the speciation 

of solid phase cadmium (e.g., Tessier et al., 1979; Hickey 

and Kittrick, 1984; Pustišek et al., 2001; Buanuam et al., 

2006). As discussed in the cited reports, sequential extraction methods provide a useful tool to assist in determining 

the chemical speciation of trace metals in soils/sediments, 

but essentially all documented methods show analytical 

limitations in selectively extracting cadmium and other 

metals associated with specific solid components. Where 

feasible, it is recommended that complimentary analytical 

techniques be employed to confirm the accuracy of cadmium speciation (e.g., O’Day et al., 1998; O’Day et al., 

2000; Carroll et al., 2002; D’Amore et al., 2005) or the 

accuracy of the extraction of a targeted phase(es) for a 

given extractant (e.g., Shannon and White, 1991; Ngiam 

and Lim, 2001; Peltier et al., 2005). As an example, Peltier 

et al. (2005) have demonstrated that a common extraction 

method employed to target metals associated with easily 

reducible iron (hydr)oxides may also dissolve iron sulfides 

that may be present. The results from this analysis may 

lead to misidentification of a cadmium association with iron 

(hydr)oxides, resulting in the development of a conceptual 

site model that misrepresents the site-specific attenuation 

process. Under reducing conditions, it is also critical that 

aquifer sediments be sampled and processed in a manner 

that prevents exposure to oxygen prior to extraction in order 

to limit oxidation of reduced minerals (e.g., iron sulfides) that 

may host cadmium. Cadmium associated with a sulfidic 

phase in sediments has been shown to repartition to more 

extractable phases upon oxidation (Saeki et al., 1993). 

Determination of the host mineral phase(es) dissolved 

for each extraction step is recommended, along with the 

use of surrogate Cd-bearing phases spiked into the sediment to confirm accuracy of the procedure (e.g., Rudd et 

al., 1988). The choice of appropriate cadmium surrogate 

phases would be governed by site-specific geochemical 

conditions or characterization of the mineralogy of the 

aquifer sediment.

Long-Term Stability and Capacity 

The stability of attenuated cadmium will largely depend on 

the stability of site-specific geochemical conditions through 

time. For example, if cadmium attenuation follows a calcium carbonate coprecipitation pathway, then long-term 

stability of attenuated cadmium will depend, in part, on 

the persistence of pH conditions. If pH conditions were to 

shift significantly to low values, cadmium might be expected 

to release from the solid phase. It is therefore important 

to understand the attenuation mechanism(s) so that geo-
6

chemical triggers for mobilization can be anticipated and 

incorporated into evaluations of long-term monitoring data. 

For any proposed and identified attenuation mechanism, 

there will exist possible scenarios whereby remobilization 

could occur (i.e., changes in pH or Eh). It will be essential to 

explore the likelihood of such changes in site geochemistry 

and the sensitivity of the attenuation pathway to changes 

in the prevailing geochemical conditions.

Quantifying the attenuation capacity (as defined in Volume 

1) will also require an understanding of the specific attenuation pathway(s). Attenuation capacity, for example, could 

be related to the extent that pH is buffered, the availability 

of sorptive sites in aquifer materials, or to the supply of 

electron donors needed to sustain microbially mediated redox conditions. For any proposed attenuation mechanism, 

there will be assumptions built into capacity estimations, 

so it is recommended that uncertainty analysis accompany 

capacity calculations.

Tiered Analysis

Determination of the viability of cadmium remediation via 

monitored natural attenuation will depend upon proper assessment of contaminant loading to the aquifer and prevailing geochemistry and mineralogy within the contaminant 

plume and the down gradient zone prior to the point(s) of 

compliance. MNA may not be appropriate as a site remedy for cadmium contamination in acidic to circum-neutral 

pH, highly oxidizing, and/or DOC-rich environments. The 

goal of site assessment is to demonstrate the process(es) 

controlling cadmium sequestration onto aquifer solids and 

the long-term stability of solid phase cadmium as a function 

of existing and anticipated ground-water chemistry. The 

following tiered analysis structure for site characterization 

provides an approach to evaluate candidate sites and define the potential limitations of MNA as part of a remedy 

for ground-water cleanup.

Tier I - Site characterization under Tier I will involve demonstration that the plume is static or shrinking, has not 

reached compliance boundaries, and does not impact 

existing water supplies. Once this is established through 

ground-water characterization, evidence is collected to 

demonstrate Cd partitioning to aquifer solids within the 

plume. Rapid movement of contaminants along preferred 

flow paths in the unsaturated and saturated zones can by 

induced by hydrologic events such as heavy rains. It will be 

important to determine that such hydrogeologic features do 

not result in contaminants bypassing zones where natural 

attenuation is occurring. If natural attenuation processes 

are active throughout the plume, then there should be an 

observed increase in solid phase concentrations within 

regions of the plume with higher aqueous concentrations, 

e.g., near the source term. This field partitioning data 

may be supplemented by geochemical modeling that incorporates measured water chemistry (e.g., pH, Eh, and 

major ion chemistry) throughout the plume to assess the 

potential for solubility control by a cadmium precipitate such 

as a carbonate/phosphate or sulfide phase. Identification 

of active sequestration to prevent cadmium migration in 

ground-water provides justification for proceeding to Tier 

II characterization efforts. 

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

of attenuation are determined. Estimates of a site attenuation rate(s) can be assessed via a well transect along 

the ground-water flow path. In addition, time-series data 

may be collected at one or more monitoring points within 

the plume. This information will allow assessment of the 

relative timescales for contaminant immobilization and fluid 

transport and determination of whether remediation objectives can be met within the required regulatory time frame. 

In addition, the mechanism(s) for attenuation need to be 

identified under this stage of site characterization. This 

effort may require determination of the chemical speciation 

of aqueous and solid phase Cd, which may be approached 

according to the following scheme:

1) Determination of cadmium solution speciation via 

direct analytical measurements in combination with 

speciation calculations based on characterized 

ground-water chemistry;

2) Calculation of the saturation state of ground water 

relative to measured aqueous chemistry complimented by the possible isolation of discrete Cd 

mineral phases via density separations (or other 

schemes) in regions of the aquifer with highest 

solid phase concentrations;

3) Determination of aquifer mineralogy to determine 

the relative abundance of components with documented capacity for Cd sorption (e.g., Amonette, 

2002);

4) Identification of cadmium association(s) with the 

various solid phase components of aquifer solids 

through combination of chemical extractions with 

microscopic/spectroscopic confirmation of phase 

associations, and;

5) Demonstration of concurrence between the site 

conceptual model and mathematical model(s) that 

describe cadmium removal mechanism(s).

It is recommended that identification of cadmium chemical 

speciation in aqueous and solid matrices be conducted 

using samples collected in a manner that preserves the 

in-situ distribution of dissolved cadmium and mineralogy 

and prevents loss of cadmium from aqueous samples 

(e.g., due to oxidation and precipitation of ferrous iron in 

anoxic ground water). The demonstration of concurrence 

between conceptual and mathematical models describing 

cadmium transport will entail development of site-specific 

parameterization of the chemical processes controlling 

cadmium solid phase partitioning.

Tier III - Once the partitioning mechanism(s) have been 

identified for the site, the subsequent characterization effort 

under Tier III will involve determination of the stability of 

immobilized Cd and the capacity of the aquifer to sustain 

continued uptake. It is recommended that the stability of 

immobilized Cd be tested based on the anticipated evolution 

of ground-water chemistry concurrent with plume shrinkage. 
7

For example, changes in ground-water pH can exert a significant influence on Cd adsorption or precipitate solubility. 

Therefore, it is recommended that sediment leach tests be 

conducted to characterize the magnitude of Cd mobilization 

as a function of pH for a ground-water chemistry representative of site conditions. It is recommended that the capacity 

for Cd uptake onto aquifer solids be determined relative to 

the specific mechanism(s) identified in Tier II. For example, 

if site characterization under Tier II indicated that precipitation of Cd sulfide due to microbial degradation of organic 

compounds coupled with sulfate reduction occurs within the 

aquifer, then it is recommended that the mass distribution 

of organic carbon and sulfate to support this reaction within 

the aquifer be determined. This site-specific capacity can 

then be compared to Cd mass loading within the plume 

in order to assess the longevity of the natural attenuation 

process. If site-specific tests demonstrate the stability of 

immobilized Cd and sufficient capacity within the aquifer to 

sustain Cd attenuation, then the site characterization effort 

can progress to Tier IV. For cases where contaminant stability is sufficient but aquifer capacity is insufficient for capture 

of the entire plume, then a determination of the benefits of 

contaminant source reduction may be necessary.

Tier IV – Finally, under Tier IV a monitoring plan is established along with contingency plans in the event of MNA 

failure. It is recommended that the monitoring plan be 

designed to establish both continued plume stability and to 

identify changes in ground-water chemistry that may lead 

to re-mobilization of attenuated Cd. The specific chemical 

parameters to be monitored will include those identified under Tier III that may halt Cd partitioning to aquifer sediments 

and/or result in solubilization of either discrete Cd precipitates or aquifer minerals that sequester Cd from ground 

water. For example, solution phase parameters that could 

alter either Cd precipitation or adsorption include increases 

in soluble organic carbon in combination with changes in 

ground-water pH. In contrast, the concentration of dissolved 

iron or sulfate may indicate the dissolution of an important 

sorptive phase within the aquifer (e.g., reductive dissolution 

of iron oxides or oxidative dissolution of sulfides). Changes 

in these parameters may occur prior to observed changes 

in solution Cd and, thus, serve as monitoring triggers for 

potential MNA failure. In this instance, a contingency plan 

can be implemented that incorporates strategies to arrest 

possible plume expansion beyond compliance boundaries. 

Possible strategies to prevent plume expansion include 

pump and treat operations, installation of reactive barriers 

to enhance uptake capacity perpendicular to the direction 

of plume advance, or enhancement of natural attenuation 

processes within the aquifer through the injection of soluble 

reactive components.

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Tessier, A., P.G.C. Campbell, and M. Bison. Sequential extraction procedure for the speciation of particulate trace 

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Tonkin, J.W., L.S. Balistrieri, and J.W. Murray. Modeling sorption of divalent metal cations on hydrous manganese 

oxide using the diffuse double layer model. Applied 

Geochemistry 19: 29-53 (2004).

USEPA. Technology Alternatives for the Remediation 

of Soils Contaminated with As, Cd, Cr, Hg, and Pb, 

EPA/540/S-97/500, U.S. Environmental Protection 

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USEPA. National Recommended Water Quality Criteria, 

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Wang, F. and A. Tessier. Cadmium complexation with 

bisulfide. Environmental Science and Technology 33: 

4270-4277 (1999).

Wang, Y. and E.J. Reardon. A siderite/limestone reactor to 

remove arsenic and cadmium from wastewaters. Applied Geochemistry 16: 1241-1249 (2001).

Wolke, R.L. Cadmium. Chemical and Engineering News

September 8, p. 120 (2003).

Zachara, J.M., C.E. Cowan, and C.T. Resch. Sorption of 

divalent metals on calcite. Geochimica et Cosmochimica Acta 55: 1549-1562 (1991).

Zhao, X., W.H. Höll, and G. Yun. Elimination of cadmium 

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11

Lead

Richard T. Wilkin, Patrick V. Brady, and Douglas B. Kent

Occurrence and Distribution

Lead is distributed in low concentrations in sedimentary 

rocks and soils. The average concentration of lead in 

shales, sandstones, and carbonate rocks is 20, 7, and 9 mg 

kg-1, respectively (Turekian and Wedepohl, 1961). KabatasPendias and Pendias (1984) report background soil concentrations of 17-26 mg Pb kg-1 in the U.S. Anthropogenic 

enrichment of lead in near-surface soils stems largely from 

airborne deposition of particles derived from fossil fuel combustion (e.g., gasoline and coal). Lead is a common metal 

contaminant at hazardous waste sites, especially at battery 

crushing and recycling facilities (USEPA, 1991). Indeed, 

lead is the most commonly recycled metal: roughly 50% of 

lead production is secondary lead. Approximately 70% of 

world-lead goes to lead-acid storage batteries. Natural lead 

enrichment occurs around hydrothermal deposits and base 

metal ores, most frequently as the mineral galena (PbS), 

but also as the oxidation products of lead sulfide ores such 

as anglesite (PbSO4) and cerussite (PbCO3).

Sources of lead contamination to surface waters and ground 

waters include: fall-out of atmospheric dust, industrial and 

municipal wastewater effluent, mineral fertilizers and pesticides, lead-based paints, and wastes from the mining, 

metallurgical, chemical, and petrochemical industries. Lead 

is a widely used non-ferrous metal in the petroleum and 

storage battery industries. In the early 20th century, lead 

was frequently used for constructing water pipes and for 

the solder used to seal pipe joints, and prior to 1978 lead 

carbonates and oxides were common pigment components 

in exterior and interior paints. From 1923 to the mid-1980’s 

tetraethyl lead was used in the U.S. as an antiknock additive in gasoline, and lead derived from fuel combustion 

represented a dominant source of lead to the environment. 

Due to a better understanding of the health consequences 

stemming from lead exposure, as well as the introduction 

of catalytic converters, many countries have reduced or 

eliminated use of lead additives in gasoline. Most of the 

lead produced in the U.S. comes from mines in Missouri, 

with other major lead-producing mines in Alaska, Colorado, 

Idaho and Montana. The average annual consumption of 

lead in the U.S. from 1997 to 2001 was 1,690,000 metric 

tons. 

Lead was ranked second on the CERCLA Priority List of 

Hazardous Substances in 1999 and 2001 (after arsenic - 

#1, and before mercury - #3, vinyl chloride - #4, and PCBs 

- #5). The priority list is prepared by the Agency for Toxic 

Substances and Disease Registry and EPA and is based 

on the frequency of occurrence of particular contaminants 

at National Priorities List (NPL) sites and their potential 

threat to human health. In absolute numbers, lead is by 

far the most common inorganic contaminant found at Superfund sites. For example, in 1996 lead contamination 

was found at 460 Superfund sites, compared to 306 with 

chromium contamination, 235 with arsenic, 226 with zinc, 

224 with cadmium, 201 with copper, and 154 with mercury 

(USEPA, 1996). 

Plume Characteristics

The fate of lead in the subsurface is controlled principally 

by adsorption at the solid-water interface, precipitation, and 

complexation with organic matter. Lead is strongly retained 

in soils and in most situations very little lead is transported 

to surface waters or ground water. Exceptions to this 

behavior are low pH systems or environments with high 

concentrations of dissolved organic carbon. Tetramethyl 

lead, a more soluble and volatile form of lead, may form as 

a result of microbial alkylation of inorganic lead in anaerobic 

environments. Remedial case studies at lead-contaminated 

Superfund sites primarily describe soil cleanup technologies (U.S. EPA, 1997). Fewer examples are found where 

remediation efforts have targeted lead contamination in 

ground water (Morrison and Spangler, 1993).

Remedial Technologies

Technology classes potentially applicable to the remediation of lead-contaminated soils include containment, 

solidification/stabilization, and separation/concentration 

(e.g., USEPA, 1997). Containment technologies applied at 

metal contamination sites include caps and vertical barriers 

to minimize the transport of lead and co-contaminants out 

of source zones. Solidification/stabilization technologies 

are treatment processes that mix reactive agents with 

contaminated material to reduce solubility or otherwise limit 

contact between the contaminated material and potential 

transport fluids. Amendments such as Portland cement 

or phosphate-based compounds are candidates for treatment of lead contamination in soils (e.g., USEPA, 1997). 

Separation/concentration methods have also been used for 

lead treatment, including ex-situ soil washing and in-situ 

soil flushing to physically or chemically reduce contaminant 

concentrations to meet site-specific cleanup goals. Groundwater remediation of lead using the permeable reactive 

barrier technology has been explored with some success 

in bench-top studies (e.g., Shokes and Möller, 1999).
12

Regulatory Aspects

Because of the highly toxic effects of lead on biological 

systems, treatment standards and concentration limits are 

stringent for the discharge of lead-bearing wastewaters 

and ground water. EPA has set the maximum contaminant 

level for lead at 0.015 mg L-1. Generally, the predominant 

routes of exposure for lead are ingestion and inhalation of 

lead-bearing aerosols. As will be discussed below, lead is 

sparingly soluble in water over a wide range of chemical 

conditions. For non-potable water sources, ambient water 

quality criteria (AWQC) that are protective of aquatic life 

may serve as alternative cleanup goals. For lead, current 

statutes list both acute and chronic criteria for fresh waters 

as 0.065 mg L-1 and 0.0025 mg L-1, respectively, for a water 

hardness of 100 mg L-1 (USEPA, 2006; http://www.epa.

gov/waterscience/criteria/nrwqc-2006.pdf). Adjustments 

to these criteria are to be applied for waters with different 

hardness. An example of where this criterion may apply 

is a site where contaminated ground water discharges to 

surface water. Cleanup goals for lead in soils at Superfund 

sites range from 200 to 500 mg kg-1 (USEPA, 1997).

Geochemistry and Attenuation Processes

Aqueous Speciation

Lead is known to form stable aqueous complexes with OH-, 

Cl-, CO3

2-, SO42-, and HS-

. In pure water, lead is mainly 

present as Pb2+ below a pH of about 7. With increasing 

pH, the species PbOH+, Pb(OH)2

0, and Pb(OH)3-

 become 

dominant over Pb2+ (Figure 2.1). Lead carbonato complexes 

(PbCO3

0, Pb(CO3)22-), lead chloride complexes (PbCl+, 

PbCl2

0), lead sulfate complexes (PbSO40, Pb(SO4)22-), 

and lead sulfide complexes (PbHS+, Pb2S2) are typically 

considered in aqueous speciation modeling efforts (e.g., 

Hem and Durum, 1973; Hem, 1976; Marani et al., 1995; 

Pierrard et al., 2002; Rozan et al., 2003). In general, complexation of lead with chloride and sulfate is most important 

at near-neutral to low pH; carbonate complexation is most 

important at near-neutral to moderately alkaline conditions; 

and, hydroxyl species dominate lead complexation at high 

pH (Figure 2.2). Inorganic speciation of lead in site-specific 

water will depend on pH, total lead concentration, and the 

relative and absolute abundances of the major anions: 

chloride, sulfate, and carbonate.

Dissolved organic carbon (DOC) may also form stable 

complexes with lead and play an important role in governing 

lead mobility in ground-water systems; however, few data 

are available and comparatively few attempts have been 

made to assess the importance of lead interactions with 

DOC. In a recent study of landfill leachate-polluted ground 

water containing up to 180 mg DOC L-1, more than 90% of 

the total lead in solution was present in DOC complexes 

(Christensen et al., 1999). This study also showed that the 

default database for MINTEQA2, which contains information 

for calculating metal complexation by DOC, was adequate 

for predicting the extent of lead complexation by DOC.

Reed et al. (1995) took advantage of lead partitioning to organic compounds in column-scale soil flushing studies. For 

soils contaminated with Pb(II) (500 mg kg-1 from Pb(NO3)2), 

PbSO4 (10,000 mg kg-1), PbCO3 (10,000 mg kg-1), and Pbnaphthalene (400 mg kg-1), they documented Pb recoveries of 100%, 100%, 100%, and 72%, respectively, using 

0.01 M EDTA as the soil-flushing solution. These results 

demonstrate the degree to which lead can be mobilized by 

organic ligands such as EDTA.

Figure 2.2 Species distribution of lead in solution with 

100 mg L-1 chloride, 100 mg L-1 sulfate, and 

100 mg L-1 total inorganic carbon, based on 

thermodynamic data in MINTEQA2 (Allison 

et al., 1990).

Solubility

An Eh-pH diagram for lead is shown in Figure 2.3. Inspection of this diagram indicates that at the specified conditions 

lead is stable in solids across the stability field of liquid 

water. At low pH and oxidizing conditions, lead sulfate 

is stable. At near-neutral to moderately alkaline pH, lead 

carbonates are stable, and at pH > 12.5 lead hydroxide is 

stable. In sulfidic environments, lead sulfide (galena) is 

stable over a wide pH range. The Pb(IV) phase, plattnerite, 

is stable at moderately alkaline to alkaline pH and at highly 

oxidizing redox potentials. Figure 2.1 Pb(II) species distribution in pure water at 

25 °C.
13

Solution pH plays a dominant role in governing lead solubility in aqueous solution. In general, the aqueous solubility of 

lead is low at near neutral to alkaline pH. Lead is expected 

to be mobile in low pH, oxidizing conditions. Hem and Durum (1973) found that at pH>7, the equilibrium solubility of 

lead was below 0.05 mg L-1 when Pb(OH)2 and PbCO3 were 

assumed to be the solubility-controlling phases. Equilibrium 

solubility of greater than 1000 mg L-1 lead was estimated 

at pH 4 in the absence of any sulfate. Lead is usually not 

a metal of concern at mining-related sites where acid mine 

drainage is produced. This is because the weathering of 

metal sulfides, in addition, to generating acidity also produces high concentrations of sulfate, which results in the 

precipitation of anglesite (Zänker et al., 2002). 

For comparison purposes, the pH-dependent solubilities of 

lead carbonate, lead sulfate, and lead sulfide are shown in 

Figure 2.4. Lead carbonate is highly insoluble at pH>8, but 

can be highly soluble below pH 6. Consequently, acidification of a soil or sediment containing lead carbonate may 

result in lead mobilization. Lead sulfate solubility is pH-independent above pH of about 2 and the concentration of 

Pb(II) in equilibrium with lead sulfate varies inversely with 

the concentration of sulfate. Lead sulfide is highly insoluble 

even at low pH (Figure 2.4).

Figure 2.3 Eh-pH diagram for lead (total Pb = 10-5 molal, total C = 10-3 molal, total S = 10-3 molal; 

all organic lead complexes are suppressed; 

activity coefficients for all species are set 

equal to 1).

Important lead-bearing mineral phases include: lead 

hydroxide (Pb(OH)2), cerussite (PbCO3), hydrocerussite 

(Pb3(CO3)2(OH)2), anglesite (PbSO4), galena (PbS), lead 

oxide (PbO), and chloropyromorphite (Pb5(PO4)3Cl) in 

phosphate-bearing systems. In addition, plumbojarosite 

(Pb0.5Fe3(SO4)2(OH)6) has been identified as an important 

secondary precipitate and lead sink in weathered mine 

wastes (e.g., Hochella et al., 1999). Thermodynamic 

data for most of these phases may be found in a variety 

of sources (e.g., see Pierrard et al., 2002, and references 

therein). In carbonate and sulfate systems the most favored 

mineral species appear to be anglesite, cerussite, and hydrocerussite (Lindsay, 1979; Marani et al., 1995).

Lead hydroxide and lead oxide, although predicted to be 

stable based on thermodynamic reasoning, seem to be 

kinetically hindered from precipitating at room temperature 

(Marani et al., 1995). In sulfate-reducing systems, galena 

precipitation is thermodynamically and kinetically favored 

over a wide range of pH and total sulfide concentrations 

(Uhler and Helz, 1984).

Marani et al. (1995) point out that the reliability of solubility 

predictions depends on the choice of the relevant solubility 

constants used in modeling studies. Unfortunately, such 

constants are wide ranging for lead. Reasonable agreement 

between solubility predictions from equilibrium modeling 

and filterable lead concentrations measured in aged soilwater systems was obtained only with a critical selection 

of solid phases in the modeling and by appreciating kinetic 

aspects of the Pb-H2O system (Marani et al., 1995).

Figure 2.4 Pb(II) activity in equilibrium with PbCO3 (at 

total inorganic carbon equal to 0.001 molal), 

PbSO4 (at total sulfate equal to 0.1 molal), 

and PbS (at total sulfide equal to 0.001 

molal). The solubility trend can be compared to the MCL for lead of 0.015 mg L-1

or an activity of ~7.2x10-8 m assuming ideal 

behavior.

Lead phosphate minerals appear to be highly insoluble 

lead-bearing phases and remediation strategies for stabilizing lead-contaminated soils have taken advantage of 

this behavior (e.g., Ruby et al., 1994; Zhang et al., 1997). 

Hydroxyapatite and sodium phosphate monobasic have 

been used as a source of soluble phosphate to amend 
14

lead-contaminated soils. Reaction between labile lead 

phases and dissolved phosphate is rapid over a wide range 

of pH and P/Pb molar ratios and results in the formation of 

insoluble chloropyromorphite (Zhang and Ryan, 1999).

Adsorption

Adsorption of trace metals, such as lead onto oxide surfaces, has been well characterized in lab-based studies 

(e.g., Hayes and Leckie, 1986). Adsorption at mineral 

surfaces results from a set of chemical reactions between 

lead and various surface sites (Dzombak and Morel, 1990). 

Most of these reactions involve the release of H+ ions, 

which accounts for the strong dependence of adsorption 

on pH. Hydrous ferric oxide (HFO) is of particular interest 

because it is found at many contaminated sites and could 

play a major role in governing the mobility of lead, other 

metals, and metalloids (e.g., Trivedi et al., 2003; Dyer et al., 

2003). Lead adsorbs more strongly onto HFO compared to 

most other divalent metal ions (Dzombak and Morel, 1990); 

the same is true for other ferric oxides, hydrous oxides, 

aluminum oxides, oxyhydroxides, clay minerals, and poorly 

ordered Fe- and Al-containing hydroxypolymer coatings on 

natural aquifer sediments (Sposito, 1984; Coston et al., 

1995; O’Reilly and Hochella, 2003).

Long-term lab studies typically indicate that adsorption 

occurs in two steps: i) rapid initial surface sorption or exchange followed by, ii) continued but slow metal uptake. 

Ainsworth et al. (1994) reported on long-term aging studies 

of lead onto HFO. Consistent with previous investigations, 

they found that the adsorption of lead onto HFO increased 

sharply from 0 to 100% as pH increased from 3 to 6. The 

pH at which 50% of the lead was associated with the 

HFO surface was 4.7. The sorption behavior of lead onto 

HFO was found to be independent of time from 0 to 21 

weeks of aging. In addition, desorption of lead from aged 

HFO exhibited no hysteresis with the adsorption pH edge 

developed from non-aged HFO. Ainsworth et al. (1994) 

concluded that the lead adsorption-desorption process is 

completely reversible with time and that there was no indication that lead was being incorporated into the HFO solid 

during aging. The process of recrystallization or aging may 

in fact result in the net loss of available sorption sites due 

to surface area reductions and may drive lead desorption 

(e.g., Ford et al., 1997).

Results of a series of laboratory experiments, which included studies of lead transport through columns packed 

with soil, were consistent with these concepts. Experiments 

showed that lead mobility decreased with increasing pH; 

lead adsorption resulted in decreases in pH, which, in 

turn, increased lead mobility; and, lead adsorption onto the 

soil was fast and reversible on the time-scale of transport 

(Papini et al., 1999).

Molecular studies of lead sorption onto hydrous ferric oxide show that Pb(II) ions associate with the iron hydroxide 

surface mainly as inner-sphere complexes (Trivedi et al., 

2003). For most of the iron oxides, edge-sharing bidentate 

complexes are dominant at pH>5 over a wide range of adsorbate concentrations (e.g., Bargar et al., 1997; Manceau 

et al., 1992). Rouff et al. (2004) reached similar conclusions 

regarding lead sorption at the calcite-water interface. However, at higher initial concentrations of lead (4-12 mg L-1), 

precipitation of lead carbonate dominates lead partitioning 

in the solid phase (Rouff et al., 2004).

Although lead adsorption in laboratory-based studies may 

be completely reversible, uptake of lead in natural systems 

is often substantially irreversible. Coughtrey et al. (1986) 

reviewed soil measurements and suggested that only 50% 

of lead in soil was exchangeable. Others have noted substantially lower exchangeable fractions (see, e.g., Wang et 

al., 1995; Brady et al., 1999). 

Adsorption of iron can influence the mobility of Pb even 

in the presence of strong complexing ligands like EDTA. 

Results of a transport experiment conducted in a mildly 

acidic, quartz-sand aquifer showed that Pb was displaced 

from EDTA complexes by Fe(III) dissolved from aquifer sediments over short transport distances (Davis et al., 2000). 

Even though Pb forms strong complexes with EDTA, strong 

adsorption of Pb at oxide surfaces enhances the thermodynamic driving force for displacement from EDTA complexes 

at mildly acidic pH values. However, decreasing solubility 

of Fe oxides with increasing pH decreases the affinity of 

the displacement reaction with increasing pH so that at pH 

values greater than 8 the reaction is unfavorable (Xue et al., 

1995; Nowack et al., 2001). Lead was also displaced from 

EDTA complexes by Zn desorbed from Zn-contaminated 

sediments (Davis et al., 2000). These types of reactions 

may limit the extent to which complexing ligands enhance 

the transport of Pb in contaminated systems.

In reducing systems, adsorption of lead to iron sulfide surfaces is possible. Jean and Bancroft (1986) investigated 

the pH-dependent adsorption behavior of Pb on several 

iron sulfide minerals and found that the pH at which 50% 

of the lead was associated with the pyrite (FeS2) surface 

was about 6.0. They suggested that the observed adsorption behavior is controlled by the hydrolysis of Pb2+ ions, 

whereby hydrolyzed species sorb directly on sulfide surface 

groups as a monolayer.

Lead adsorption onto or co-precipitation with amorphous 

FeS may be extensive in reducing systems. Experimental 

studies suggest that a significant fraction of the Fe in freshly 

precipitated FeS may be replaced rapidly by metal ions like 

Pb that form less soluble sulfides (Phillips and Kraus, 1965; 

Caletka et al., 1975; Coles et al., 2000).

Redox Chemistry

In natural systems lead is present in the +2 oxidation state 

over relevant conditions of pH and oxidation-reduction potential. Reduction of Pb2+ to metallic lead is expected to 

occur at redox potentials below the stability field of water at 

pH<6. The oxidized form of lead (Pb(IV)) is not expected 

in air-saturated solutions based on thermodynamic reasoning. However, the mineral plattnerite (PbO2) occurs in 

some natural systems and is associated with other oxidation products such as cerussite and pyromorphite (e.g., 

Yeates and Ayres, 1892; see Figure 2.3). In general, the 

geochemical transport processes of lead are not directly 
15

tied to redox conditions. However, because lead may form 

stable precipitates with redox-sensitive elements such as 

sulfur, lead mobility is indirectly tied to redox conditions. In 

sulfate-reducing systems, lead is expected to form insoluble 

PbS precipitates. In moderately reducing but non-sulfidic 

systems, however, reductive dissolution of hydrous ferric 

oxides that contain adsorbed lead could result in lead 

mobilization.

Colloidal Transport

The transport of lead in particulate forms can be significant 

in ground water and surface water systems. Colloids are 

generally considered to be particles with diameters less 

than 10 micrometers (Stumm and Morgan, 1996). Colloidal 

particles can be present as mineral or organic forms. For 

the special case of lead sorbed to colloidal hydrous ferric 

oxides, changes in geochemical regimes may either favor 

increased lead sorption or desorption. Increases in lead 

sorption may result from increases in pH or Eh. Conversely, 

decreases in pH or Eh may result in lead remobilization. 

Decreases in the ionic strength of the aqueous phase can 

enhance colloidal stability and promote lead transport, 

whereas increases in ionic strength can promote colloid 

aggregation and removal from the aqueous phase.

Site Characterization

Lead mobility in ground water is governed by the total concentration of lead, the distribution of lead species in water, 

and the nature of lead partitioning in the solid phase. The 

development of site-conceptual models for predicting the 

long-term fate of lead at a contaminated site will require 

information on the distribution and concentration of lead in 

the aqueous phase and the solid phase. Table 2.1 indicates 

possible natural attenuation and mobilization pathways 

for lead. Details of the types of analytical measurements 

that may be conducted on sampled ground water and 

aquifer sediments to assist in identifying the attenuation 

mechanism(s) are discussed in the following paragraphs.

Aqueous Measurements

Quantitative measurement of lead concentrations in aqueous solutions is typically carried out using inductively 

coupled plasma atomic emission spectroscopy (ICP-AES), 

inductively coupled plasma mass spectroscopy (ICP-MS), 

or atomic absorption spectroscopy (AAS). Input data to 

geochemical codes (e.g., MINTEQA2, PHREEQC, EQ3/6) 

for determining aqueous speciation also require, at a 

minimum, concentrations of major anions, major cations, 

dissolved organic carbon, temperature, and pH. In addition, while regulatory requirements stipulate that unfiltered 

ground-water samples be analyzed to support regulatory 

decisions at a contaminated site, it may be necessary to 

also collect filtered samples to help define the process(es) 

controlling contaminant mobility. The use of 0.45 μm pore 

size filters is common as an arbitrary cutoff point to differentiate between dissolved and particulate phases in 

water samples. However, caution is recommended when 

using this approach, particularly for Fe and Al and other 

elements that may be associated with Fe or Al particles 

(including Pb) that could pass through 0.45 μm filters. The 

use of filters with pore sizes less than 0.1 μm will generally 

provide a better assessment of the dissolved vs. particulate 

load in ground water.

Solid Phase Measurements

The implementation of an analytical approach to identify 

lead speciation in aquifer sediments is a challenging process. The accuracy of the analytical finding is dependent 

on the method of sample collection/preservation and the 

tools used to identify the mechanism of lead partitioning. It 

is recommended that the analytical protocol be designed to 

address the potential redox sensitivity of the solid phase(s) 

to which lead may be partitioned (e.g., sulfides in reduced 

sediments). Tools to evaluate the mechanism of lead 

solid phase partitioning range in complexity from relatively 

simple chemical extractions to advanced spectroscopic 

techniques.

Table 2.1 Natural attenuation and mobilization pathways for lead.

Attenuation Processes Mobilization Processes Characterization Approach

Precipitation of insoluble carbonates, sulfides, sulfates, and 

phosphates. In general, pH>8 

will drive precipitation reactions 

resulting in Pb concentrations 

to below the MCL.

Dissolution of carbonates at 

low pH; oxidative dissolution of 

sulfides at low pH and high Eh; 

complexation/stabilization in the 

presence of DOC.

Evaluation of lead speciation in the aqueous 

phase. Determination of total Pb in the solid 

matrix. Evaluation of solid phase partitioning using sequential extraction methodologies. 

Evaluation of long-term capacity.

Sorption to iron hydroxides 

(reversible), organic matter, 

carbonates, sulfides (pH>5).

Desorption at low pH; complexation/stabilization in the 

presence of DOC. Reductive 

dissolution of iron hydroxides.

Evaluation of lead speciation in the aqueous 

phase. Determination of total Pb in the solid 

matrix. Evaluation of solid phase partitioning using sequential extraction methodologies. 

Batch and column testing to determine Pb uptake capacity of site-specific aquifer materials 

with variable geochemical conditions.
16

The total concentration of lead in soils, sediments, and 

aquifer materials may be determined by X-ray fluorescence 

spectroscopy (XRF), or by ICP-AES after digestion in 

mineral acids. A variety of digestion or extraction methods 

can be found in the literature (Amacher, 1996). Neutron 

activation analysis is not commonly employed due to the 

scarcity of neutron sources required to irradiate the sample. 

X-ray fluorescence is the most attractive approach due 

to the relative ease of sample preparation, which may be 

conducted with the sample in its original state or following 

fusion with lithium metaborate. When combined with the 

determination of other major or trace elements in the solid 

sample, this provides an initial step for assessing possible 

association of lead with various solid phase components. 

This type of analysis can be conducted on the bulk sample 

as well as at a microscopic level using wavelength (electron 

microprobe) or energy dispersive spectroscopy coupled to a 

scanning or transmission electron microscope. Microscopic 

examination allows one to better differentiate whether lead 

may be distributed across a number of different mineral 

phases within the solid sample or primarily associated with 

a discrete phase. There are limitations to this approach 

(Pye, 2004), a significant one being that the analysis does 

not necessarily provide unique mineral identification necessitating the collection of supporting mineralogical and 

chemical data.

More detailed information on the specific partitioning 

mechanism(s) controlling lead solid phase speciation is 

typically required to adequately support site assessment 

for potential reliance on natural attenuation as part of a site 

remedy. There have been many applications of sequential 

extraction schemes to assess the speciation of solid phase 

lead (e.g., Tessier et al., 1979; Harrington et al., 1998; 

Sutherland, 2002; Zänker et al., 2002; Buanuam et al., 

2006). As discussed in the cited reports, sequential extraction methods provide a useful tool to assist in determining 

the chemical speciation of trace metals in soils/sediments, 

but essentially all documented methods show analytical 

limitations in selectively extracting lead and other metals 

associated with specific solid components (e.g., Scheckel et 

al., 2003). Design and application of extraction procedures 

should take into account the chemical behavior of lead 

relative to potential analytical bias that may be introduced 

by the extraction chemistry. For example, due to its low 

solubility, lead carbonate would be anticipated to form during extractions conducted using solutions with pH buffered 

by an excess of dissolved bicarbonate-carbonate. 

Where feasible, it is recommended that complimentary 

analytical techniques be employed to confirm the accuracy of lead speciation (e.g., O’Day et al., 1998; O’Day et 

al., 2000; Carroll et al., 2002; D’Amore et al., 2005) or the 

accuracy of the extraction of a targeted phase(es) for a 

given extractant (e.g., Shannon and White, 1991; Ngiam 

and Lim, 2001; Peltier et al., 2005). As an example, Peltier 

et al. (2005) have demonstrated that a common extraction 

method employed to target metals associated with easily 

reducible iron (hydr)oxides may also dissolve iron sulfides 

that may be present. The results from this analysis may 

lead to misidentification of a lead association with iron 

(hydr)oxides, resulting in the development of a conceptual 

site model that misrepresents the site-specific attenuation process. Under reducing conditions, it is also critical 

that aquifer sediments be sampled and processed in a 

manner that prevents exposure to oxygen prior to extraction in order to limit oxidation of reduced minerals (e.g., 

iron sulfides) that may host lead. Lead associated with a 

sulfidic phase in sediments has been shown to repartition 

to more extractable phases upon oxidation (Saeki et al., 

1993; Cauwenberg and Maes, 1997). Determination of the 

host mineral phase(es) dissolved for each extraction step 

is recommended, along with the use of surrogate Pb-bearing phases spiked into the sediment to confirm accuracy 

of the procedure (e.g., Rudd et al., 1988). The choice of 

appropriate lead surrogate phases would be governed by 

site-specific geochemical conditions or characterization of 

the mineralogy of the aquifer sediment.

Pb Isotopes

Information about the source of lead contamination at a 

given site can be gained using isotopic analysis, particularly by examining the ratios 206Pb/204Pb, 207Pb/204Pb, and 208Pb/204Pb (Emmanuel and Erel, 2002). As an example, 

Chow and Johnstone (1965) reported isotope ratios of lead 

extracted from gasoline (purchased in 1965). Their data 

demonstrated the similarity of the isotope compositions of 

lead in gasoline, airborne particles in Los Angeles, and 

snow from Lassen Volcanic Park. The study showed that 

lead in the air and snow in California in the 1960’s originated 

from fuel combustion exhaust. Gulson et al. (1981) analyzed 

the lead isotopic composition of soils in South Australia to 

identify the source of lead contamination. By analyzing 

and comparing isotope ratios they determined that orchard 

sprays, power stations, and smelters were not the principal 

source of lead contamination, rather lead contamination 

was again derived from tetraethyl lead in gasoline.

Long-term Stability and Capacity

The stability of attenuated lead will depend on the temporal 

stability of site geochemical conditions. For example, if lead 

attenuation follows a lead sulfide precipitation pathway, then 

long-term stability of attenuated lead may depend on the 

persistence of reducing conditions. It is therefore critical to 

understand attenuation mechanism(s) so that geochemical 

triggers for remobilization can be anticipated and incorporated into evaluations of long-term monitoring data. For any 

proposed and identified attenuation mechanism, there will 

exist possible scenarios whereby remobilization could occur 

(i.e., changes in pH or Eh). It will be essential to explore the 

likelihood of such changes in prevailing site geochemistry 

and the sensitivity of the attenuation pathway to changes 

in the prevailing geochemical conditions.

Quantifying the attenuation capacity (as defined in Volume 

1) will also require an understanding of the specific attenuation pathway(s). Attenuation capacity, for example, could 

be related to the extent that pH is buffered, the availability 

of sorptive sites in aquifer materials, or to the supply of 

electron donors needed to sustain microbially mediated redox conditions. For any proposed attenuation mechanism, 
17

there will be assumptions built into capacity estimations, 

so it is recommended that uncertainty analysis accompany 

capacity estimates. 

Tiered Analysis

Determination of the viability of lead remediation in ground 

water via monitored natural attenuation will depend upon 

proper assessment of contaminant loading to the aquifer 

and prevailing geochemistry and mineralogy within the 

contaminant plume and the down gradient zone prior to 

the point(s) of compliance. MNA may not be appropriate as a site remedy for lead contamination in acidic to 

circum-neutral pH, highly oxidizing, and/or DOC-rich environments. The goal of site assessment is to demonstrate 

the process(es) controlling lead sequestration onto aquifer 

solids and the long-term stability of solid phase lead as a 

function of existing and anticipated ground-water chemistry. 

The following tiered analysis structure for site characterization provides an approach to evaluate candidate sites and 

define the potential limitations of MNA as part of a remedy 

for ground-water cleanup.

Tier I - Site characterization under Tier I will involve demonstration that the plume is static or shrinking, has not 

reached compliance boundaries, and does not impact 

existing water supplies. Once this is established through 

ground-water characterization, evidence is collected to 

demonstrate Pb partitioning to aquifer solids within the 

plume. Rapid movement of contaminants along preferred 

flow paths in the unsaturated and saturated zones can by 

induced by hydrologic events such as heavy rains. It will be 

important to determine that such hydrogeologic features do 

not result in contaminants bypassing zones where natural 

attenuation is occurring. If natural attenuation processes 

are active throughout the plume, then there should be an 

observed increase in solid phase concentrations within 

regions of the plume with higher aqueous concentrations, 

e.g., near the source term. This field partitioning data 

may be supplemented by geochemical modeling that incorporates measured water chemistry (e.g., pH, Eh, and 

major ion chemistry) throughout the plume to assess the 

potential for solubility control by a lead precipitate such 

as a carbonate/phosphate or sulfide phase. Identification of active sequestration to prevent lead migration in 

ground-water provides justification for proceeding to Tier 

II characterization efforts. 

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

of attenuation are determined. Estimates of a site attenuation rate(s) can be assessed via a well transect along 

the ground-water flow path. In addition, time-series data 

may be collected at one or more monitoring points within 

the plume. This information will allow assessment of the 

relative timescales for contaminant immobilization and fluid 

transport and determination of whether remediation objectives can be met within the required regulatory time frame. 

In addition, the mechanism(s) for attenuation need to be 

identified under this stage of site characterization. This 

effort may require determination of the chemical speciation 

of aqueous and solid phase Pb, which may be approached 

according to the following scheme:

1) Determination of lead solution speciation via direct analytical measurements in combination with 

speciation calculations based on characterized 

ground-water chemistry;

2) Calculation of the saturation state of ground water 

relative to measured aqueous chemistry complimented by the possible isolation of discrete Pb 

mineral phases via density separations (or other 

schemes) in regions of the aquifer with highest 

solid phase concentrations;

3) Determination of aquifer mineralogy to determine 

the relative abundance of components with documented capacity for Pb sorption (e.g., Amonette, 

2002);

4) Identification of lead association(s) with the 

various solid phase components of aquifer solids 

through combination of chemical extractions with 

microscopic/spectroscopic confirmation of phase 

associations, and;

5) Demonstration of concurrence between the site 

conceptual model and mathematical model(s) that 

describe lead removal mechanism(s).

It is recommended that identification of lead chemical speciation in aqueous and solid matrices be conducted using 

samples collected in a manner that preserves the in-situ 

distribution of dissolved lead and mineralogy and prevents 

loss of lead from aqueous samples (e.g., due to oxidation 

and precipitation of ferrous iron in anoxic ground water). 

The demonstration of concurrence between conceptual and 

mathematical models describing lead transport will entail 

development of site-specific parameterization of the chemical processes controlling lead solid phase partitioning.

Tier III - Once the partitioning mechanism(s) have been 

identified for the site, the subsequent characterization effort 

under Tier III will involve determination of the stability of 

immobilized Pb and the capacity of the aquifer to sustain 

continued uptake. It is recommended that the stability of 

immobilized Pb be tested based on the anticipated evolution 

of ground-water chemistry concurrent with plume shrinkage. 

For example, changes in ground-water pH can exert a significant influence on Pb adsorption or precipitate solubility. 

Therefore, it is recommended that sediment leach tests be 

conducted to characterize the magnitude of Pb mobilization 

as a function of pH for a ground-water chemistry representative of site conditions. It is recommended that the capacity 

for Pb uptake onto aquifer solids be determined relative to 

the specific mechanism(s) identified in Tier II. For example, 

if site characterization under Tier II indicated that precipitation of Pb sulfide due to microbial degradation of organic 

compounds coupled with sulfate reduction occurs within the 

aquifer, then it is recommended that the mass distribution 

of organic carbon and sulfate to support this reaction within 

the aquifer be determined. This site-specific capacity can 

then be compared to Pb mass loading within the plume 

in order to assess the longevity of the natural attenuation 

process. If site-specific tests demonstrate the stability of 

immobilized Pb and sufficient capacity within the aquifer to 

sustain Pb attenuation, then the site characterization effort 
18

can progress to Tier IV. For cases where contaminant stability is sufficient but aquifer capacity is insufficient for capture 

of the entire plume, then a determination of the benefits of 

contaminant source reduction may be necessary.

Tier IV – Finally, under Tier IV a monitoring plan is established along with contingency plans in the event of MNA 

failure. It is recommended that the monitoring plan be 

designed to establish both continued plume stability and to 

identify changes in ground-water chemistry that may lead 

to re-mobilization of attenuated Pb. The specific chemical 

parameters to be monitored will include those identified under Tier III that may halt Pb partitioning to aquifer sediments 

and/or result in solubilization of either discrete Pb precipitates or aquifer minerals that sequester Pb from ground 

water. For example, solution phase parameters that could 

alter either Pb precipitation or adsorption include increases 

in soluble organic carbon in combination with changes in 

ground-water pH. In contrast, the concentration of dissolved 

iron or sulfate may indicate the dissolution of an important 

sorptive phase within the aquifer (e.g., reductive dissolution 

of iron oxides or oxidative dissolution of sulfides). Changes 

in these parameters may occur prior to observed changes 

in solution Pb and, thus, serve as monitoring triggers for 

potential MNA failure. In this instance, a contingency plan 

can be implemented that incorporates strategies to arrest 

possible plume expansion beyond compliance boundaries. 

Possible strategies to prevent plume expansion include 

pump and treat operations, installation of reactive barriers 

to enhance uptake capacity perpendicular to the direction 

of plume advance, or enhancement of natural attenuation 

processes within the aquifer through the injection of soluble 

reactive components.

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21

Nickel

Kirk G. Scheckel, Robert G. Ford, Richard T. Wilkin

Occurrence and Distribution

Industrial activity and natural environmental conditions have 

led to the introduction of nickel into soil and aquatic environments as a result of anthropogenic and geogenic sources, 

respectively (Duke, 1980; Richter and Theis, 1980). Nickel 

is a relatively minor constituent of the earth’s crust having 

an average concentration of less than 0.01% by weight 

and ranking 24th in terms of abundance. Nickel is very 

heterogeneously distributed among crustal rocks ranging 

from less than 0.0001% in sandstone and granite to 4% in 

coveted ore deposits (Duke, 1980). Nickel can be found 

in igneous, sedimentary, and metamorphic rocks as well 

as nickel ores. In soils, nickel ranges from 5 – 500 mg kg-1

(Lindsay, 1979). Serpentine clay-rich soils are noted for 

natural geogenic abundance of nickel and have been the 

focus for use of hyperaccumulating plants to phytomine 

nickel (Chaney et al., 1995).

Nickel is one of the most mobile of the heavy metals in the 

aquatic environment. The mobility of nickel in the aquatic 

environment is controlled largely by competition between 

various sorbents to scavenge it from solution and ligands to 

form non-sorptive complexes. Although data are limited, it 

appears that in pristine environments, hydrous oxides and 

phyllosilicates control nickel mobility via co-precipitation 

and sorption. In polluted environments, the more prevalent organic compounds will keep nickel soluble by ligand 

complexation. In reducing environments, insoluble nickel 

sulfide may form. Nickel chloride is water-soluble and 

would be expected to release divalent nickel into the water. 

The atmosphere is a major conduit for nickel as particulate 

matter. Contributions to atmospheric loading come from 

both natural sources and anthropogenic activity, with input 

from both stationary and mobile sources. Various dry and 

wet precipitation processes remove particulate matter as 

wash out or fallout from the atmosphere with transfer to 

soils and waters. Soil borne nickel may enter waters by 

surface runoff or by percolation into ground water. Once 

nickel is in surface and ground-water systems, physical 

and chemical interactions (complexation, precipitation/dissolution, adsorption/desorption, and oxidation/reduction) 

occur that will determine its fate and that of its constituents. 

The only gaseous nickel compound of environmental importance is nickel carbonyl. Under ambient conditions in 

moist air, it decomposes to form nickel carbonate. Thus, 

in the atmosphere at concentrations near the ppb level, it 

has a half-life of about 30 minutes. The removal of nickel 

carbonyl by precipitation or by adsorption on surfaces has 

not been documented. Since this compound is soluble 

in water, precipitation scavenging is possible. Nothing 

is known about its reaction with natural surfaces or its 

uptake by vegetation. Thus, dry deposition rates cannot 

be predicted until some experimental investigations have 

been conducted. Although nickel is bioaccumulated, the 

concentration factors are such as to suggest that partitioning into the biota is not a dominant fate process.

Production of nickel was 84.6 million pounds in 1986, 

down slightly from 90 million pounds reported in 1982. In 

1986 it was estimated that industries consumed nickel as 

follows: transportation, 25%, chemical industry, 15%; electrical equipment, 9%; construction, 9%; fabricated metal 

products, 9%; petroleum, 8%; household appliances, 7%; 

machinery, 7%; and other, 11%. Nickel carbonate is used 

in nickel catalyst production for organic chemical manufacture, petroleum refining and edible oil hardening. Nickel 

oxide consumption in 1972 (representing over 30 million 

pounds containing nickel) is estimated to have been as 

follows: 60% for stainless and heat resisting steels, 27% 

for other steel alloys, 8% for other nickel alloys, 2% for cast 

irons, and 3% for other uses (USEPA, 1986). From 1987 

to 1993, according to the Toxics Release Inventory nickel 

released to land and water totaled nearly 27 million pounds, 

of which most was to land. These releases were primarily 

from nickel smelting/refining and steelworks industries. The 

largest releases occurred in Oregon and Arkansas. The 

largest direct releases to water occurred in Maryland and 

Georgia (USEPA, 2003).

Plume Characteristics

The mobility of nickel in ground water will be controlled 

by partitioning reactions to aquifer sediments. Possible 

mechanisms influencing nickel partitioning to subsurface 

solids include direct adsorption to clay minerals, adsorption and/or coprecipitation with metal oxides, complexation 

with natural organic particles, ion exchange with charged 

surfaces, and direct precipitation as an hydroxide, carbonate or sulfide (Snodgrass, 1980). The chemical speciation 

of nickel in solution exerts a significant influence on the 

extent and mechanism(s) of partitioning to aquifer sediments, which may be influenced by acid-base reactions, 

oxidation-reduction reactions influencing the speciation 

of complexing inorganic solution species (e.g., aqueous 

sulfate vs. sulfide), and interactions with dissolved organic 

compounds. In general, inorganic/organic species that form 

dissolved complexes with nickel tend to enhance transport 
22

of nickel in soil profiles to subsurface water (e.g., dissolved 

organic carbon; Christensen et al., 1996; Warwick et al., 

1997; Christensen and Christensen, 2000; Friedly et al., 

2002). Field studies on transport in the subsurface illustrate 

several general conditions that are anticipated to result in 

expanding nickel plumes, including 1) acidic conditions 

(Kjoller et al., 2004), 2) manganese- and iron-reducing 

conditions (Larsen and Postma, 1997), and 3) the presence of mobile organic compounds that form soluble nickel 

complexes (Christensen et al. 1996; Kent et al., 2002). 

Remedial Technologies

Possible engineered approaches that can be employed 

for remediation of a ground-water plume containing nickel 

include physical removal of contaminated soils or sediments 

that serve as a long-term source of nickel leached into 

ground water, extraction of the dissolved plume with some 

method of above-ground treatment, physical isolation of the 

dissolved plume, or in-situ treatment of a dissolved plume 

resulting in immobilization of dissolved nickel within the 

aquifer. Of these technologies, the use of permeable reactive barriers (PRBs) for the capture and immobilization of 

nickel plumes has been investigated and applied in field settings due to favorable performance and cost characteristics 

(Blowes et al., 2000). Both carbon- and metallic iron-based 

(or zero valent iron) reactive media have been employed for 

nickel removal from ground water. For carbon-based media, 

nickel removal is generally considered to occur through the 

precipitation of sulfide minerals, including nickel sulfides or 

coprecipitation of nickel with iron sulfides (e.g., Ludwig et al., 

2002; McGregor et al., 2002). Reactive sulfide is generated 

in this type of PRB as a result of microbial sulfate reduction 

stimulated by degradation of an organic carbon substrate 

incorporated into the reactive barrier media. Zerovalent 

iron media have also been tested for the removal of nickel 

in ground water (e.g., Wilkin and McNeil, 2003). For this 

material, nickel removal may be achieved either through the 

stimulation of sulfate reduction with precipitation as a sulfide 

or through coprecipitation with or adsorption onto metallic 

iron corrosion products such as iron (hydr)oxides. There 

is also laboratory and field evidence that nickel immobilization can be enhanced through the addition of chemical 

amendments that promote nickel precipitation within soil or 

aquifer sediments (e.g., Lothenbach et al., 1997; Boisson 

et al., 1999; Seaman et al., 2001). The applicability and 

performance of these technologies will depend on the geochemical characteristics within the ground-water plume in 

conjunction with the velocities of ground-water flow and the 

flux of beneficial and non-beneficial reactive components 

transported within the plume.

Regulatory Aspects

In 2005, nickel was ranked 55 of 275 hazardous substances 

on the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA) National Priorities 

List (NPL) based on frequency of occurrence at NPL sites, 

toxicity, and potential for human exposure to the substances 

found at NPL sites (ATSDR, 2005; http://www.atsdr.cdc.

gov/cercla/supportdocs/text.pdf). There are currently no 

primary or secondary drinking water standards (maximum 

contaminant level or MCL) in place for nickel in potable 

water sources (USEPA, 2006a; See also http://www.epa.

gov/safewater/dwh/t-ioc/nickel.html). However, the health 

advisory for nickel, an estimate of acceptable drinking water 

levels for a chemical substance based on health effects information, via consumption of water has been set at 1 mg L-1

for one- and ten-day exposures for a child and 0.1 mg L-1

for a lifetime exposure for an adult (USEPA, 2006a; http://

www.epa.gov/waterscience/criteria/drinking/dwstandards.

pdf). For non-potable water sources, ambient water quality criteria (AWQC) that are protective of aquatic life may 

serve as alternative cleanup goals. For nickel, current 

statutes list both acute and chronic criteria for fresh waters 

as 0.47 mg L-1 and 0.052 mg L-1, respectively, for a water 

hardness of 100 mg L-1 (USEPA, 2006b; http://www.epa.

gov/waterscience/criteria/nrwqc-2006.pdf). Adjustments 

to these criteria are to be applied for waters with different 

hardness. An example of where this criterion may apply 

is a site where contaminated ground water discharges to 

surface water. 

Geochemistry and Attenuation Processes

Aqueous Speciation

In ambient aqueous systems, nickel exists in the divalent 

oxidation state and is not subject to oxidation-state transformations under typical conditions. Nickel predominantly 

exists as a cationic species (Ni2+) or various hydolysis species (e.g., NiOH+) at near-neutral pH (Baes and Mesmer, 

1986). However, nickel may also form dissolved complexes 

in the presence of high concentrations of inorganic ions 

such as carbonate/bicarbonate and sulfate (Hummel and 

Curti, 2003; Chen et al., 2005) or organic ligands such 

as natural/synthetic carboxylic acids and dissolved humic 

compounds (Bryce and Clark, 1996; Baeyens et al., 2003; 

Strathman and Myneni, 2004). It is anticipated that nickel 

may form complexes with dissolved sulfide under sulfatereducing conditions, although the current state of knowledge 

is insufficient to ascertain the relative importance of these 

species in aqueous systems (Thoenen, 1999). The formation of solution complexes, especially with organic ligands, 

may limit sorption of nickel to mineral surfaces in aquifer 

sediments (see Adsorption section below). 

Solubility

Nickel may be immobilized within ground water through 

formation of pure nickel precipitates such as hydroxides, 

silicates, or sulfides (Merlen et al., 1995; Mattigod et al., 

1997; Scheidegger et al., 1997; Thoenen, 1999; Scheinost 

and Sparks, 2000; Peltier et al., 2006) or through coprecipitation with other soil forming minerals such as silicates, 

iron oxides/sulfides, or carbonates (Manceau et al., 1985; 

Manceau and Calas, 1986; Huerta-Diaz and Morse, 1992; 

Ford et al., 1999a; Hoffmann and Stipp, 2001). Predicted 

nickel concentrations in the absence of sulfide for several 

potential pure nickel precipitates are shown in Figure 3.1. 

These data suggest that phyllosilicate and layered double 

hydroxide (LDH) precipitates (incorporating aluminum) may 

result in dissolved nickel concentrations below most relevant 

regulatory criteria over a pH range typical for ground water. 
23

These data also point to the limited capability of pure nickel 

carbonates and hydroxides in controlling dissolved nickel 

concentrations to sufficiently low values except under very 

alkaline conditions. In the presence of dissolved sulfide, 

the precipitation of a nickel sulfide may plausibly control the 

concentration of dissolved nickel. The Eh-pH conditions 

under which these solubility-limiting phases may form is 

shown in Figure 3.2. According to these data, nickel-bearing phyllosilicate and/or LDH precipitates possess large 

Figure 3.1 (a) Predicted solubility of various Ni precipitates that could form in aerobic ground water with concentrations of Al and Si controlled by the solubility of the clay mineral, kaolinite. (Note y-axis break to expand 

lower end of scale.) (b) Expansion of dissolved Ni data for equilibrium with Ni phyllosilicate; plotted on 

logarithmic scale. Nominal ground-water composition: 0.005 mole/L NaCl, 0.001 mole/L K2SO4, 0.001 

mole/L MgNO3, 0.001 mole/L CaCO3, and 34 μ mole/L Ni (2000 μg Ni/L). Model predictions using Visual 

MINTEQ Version 2.50 (Based on MINTEQA2 described in Allison et al. (1991); available at http://www.

lwr.kth.se/English/OurSoftware/vminteq/) with solubility constants added for Ni phyllosilicate, Ni-Al-SO4

LDH, and Ni-Al-CO3 LDH (Peltier et al., 2006); total dissolved Ni concentrations modeled individually for 

each Ni solid phase by setting kaolinite as an ‘infinite’ solid and the Ni solid phase as a ‘possible’ solid for 

each pH titration.

Figure 3.2 Eh-pH diagrams for nickel at 25 °C. (a) System Ni-H2O-Ca-Al-NO3-HCO3-SO4 (2 mg Ni/L; 40 mg Ca/L; 

3 mg Al/L; 6 mg NO3/L; 60 mg HCO3/L; 100 mg SO4/L). Stability fields for solids are shaded green 

(Vaesite = NiS2). (b) Same system plus 3 mg Si /L. Thermodynamic data for Ni3Si4O10(OH)2 and 

Ni0.63Al0.33(OH)2(SO4)0.125 are from Peltier et al. (2006). [Note that the solubility of the Ni-Al-SO4 LDH was 

adjusted to correct for charge imbalance for the chemical structure published in Peltier et al. (2006).] 

stability fields indicating their relative importance to controlling nickel solubility under a range of conditions. These 

calculations point to the importance of dissolved aluminum 

and silicon concentrations in ground water relative to the 

potential sequestration of nickel via precipitation (Ford et al., 

1999b; Scheinost et al., 1999). As discussed below (see 

Adsorption section), the formation of these nickel-bearing 

precipitates may be facilitated through initial adsorption 

onto clay minerals within the aquifer.
24

Attenuation of nickel may also occur via coprecipitation during the formation of (hydr)oxides or sulfides of 

iron. These minerals have been observed to form at the 

boundaries between oxidizing and reducing zones within 

ground-water plumes. There are numerous laboratory and 

field observations that demonstrate the capacity of these 

precipitates for nickel uptake (Schultz et al., 1987; HuartaDiaz and Morse, 1992; Coughlin and Stone, 1995; Ford et 

al., 1997; Ford et al., 1999a). Under these circumstances, 

the solubility of nickel will depend on the stability of the host 

precipitate phase. For example, iron oxide precipitates 

may alternatively transform to more stable forms (Ford et 

al., 1997), stabilizing coprecipitated nickel over the long 

term, or these precipitates may dissolve concurrent with 

changes in ground-water redox chemistry (e.g., Zachara 

et al., 2001).

Adsorption

Adsorption of nickel in soil environments is dependent on 

pH, temperature, and type of sorbent (minerals or organic 

matter), as well as the concentration of aqueous complexing agents, competition from other adsorbing cations, and 

the ionic strength in ground water. Nickel has been shown 

to adsorb onto many solid components encountered in 

aquifer sediments, including iron/manganese oxides, clay 

minerals (Dähn et al., 2003; Bradbury and Baeyens, 2005), 

and solid organic matter (Nachtegaal and Sparks, 2003). 

Sorption to iron/manganese oxides and clay minerals has 

been shown to be of particular importance for controlling 

nickel mobility in subsurface systems. The relative affinity 

of these individual minerals for nickel uptake will depend 

on the mass distribution of the sorbent minerals as well 

as the predominant geochemical conditions (e.g., pH and 

nickel aqueous speciation). For example, the pH-dependent distribution of nickel between iron and manganese 

oxides [hydrous ferric oxide (HFO) and a birnessite-like 

mineral (nominally MnO2)] for a representative groundwater composition is shown in Figure 3.3a. Based on the 

available compilations for surface complexation constants 

onto these two solid phases (Dzombak and Morel, 1990; 

Tonkin et al., 2004), one would project the predominance 

of nickel sorption to MnO2 at more acidic pH and the predominance of HFO (or ferrihydrite) at more basic pH. With 

increasing mass of MnO2, the solid-phase speciation of 

nickel will be progressively dominated by sorption to this 

phase. There are examples of the relative preference of 

nickel sorption to manganese oxides over iron oxides for 

natural systems (e.g., Larsen and Postma, 1997; Manceau 

et al., 2002; Kjoller et al., 2004; Manceau et al., 2006). As 

shown in Figure 3.3b, nickel adsorption may be inhibited 

(or nickel desorption enhanced) through the formation of 

solution complexes with organic ligands such as EDTA 

or natural organic matter (e.g., Bryce and Clark, 1996; 

Nowack et al., 1997). These dissolved compounds may be 

present as natural components within ground water or as 

co-contaminants within a contaminant plume (e.g., Means 

et al., 1978).

Figure 3.3 (a) Nickel sorption as a function of pH in the presence of an hypothetical aquifer sediment with iron and 

manganese oxides reflective of the crustal abundance of these elements (Schulze, 2002; assumed 30% 

porosity with 185.0 g HFO/L and 1.66 g MnO2/L). (b) Same conditions as in (a), but with 10 μM EDTA 

added. Nominal ground-water composition: 0.005 mole/L NaCl, 0.001 mole/L K2SO4, 0.001 mole/L 

MgNO3, 0.001 mole/L CaCO3, and 34 μ mole/L Ni (2 mg Ni/L). Model predictions using Visual MINTEQ 

Version 2.50 (Based on MINTEQA2 described in Allison et al. (1991); available at http://www.lwr.kth.

se/English/OurSoftware/vminteq/) with available surface complexation parameters derived from Dzombak 

and Morel (1990) and Tonkin et al. (2004); kaolinite set as an ‘infinite’ solid for pH titration.
25

As previously noted, adsorption of nickel onto mineral 

surfaces may serve as a precursory step to the formation 

of trace precipitates that reduce the potential for desorption with changes in ground-water chemistry. This may 

be realized through the nucleation and growth of surface 

precipitates on clay mineral surfaces due to continued 

uptake of nickel (Scheckel and Sparks, 2000; Scheckel et 

al., 2000; Scheckel and Sparks, 2001; Dähn et al., 2002). 

This type of process may compete with other adsorption 

processes, such as ion exchange, depending on the prevailing ground-water chemistry and characteristics of the 

clay mineral (Elzinga and Sparks, 2001). 

Site Characterization

Overview

Nickel mobility in ground water is governed by the total 

concentration of nickel, the distribution of nickel species 

in water, and the nature of nickel partitioning in the solid 

phase. The development of site-conceptual models for 

predicting the long-term fate of nickel at a contaminated 

site will require information on the distribution and concentration of nickel in the aqueous phase and the solid 

phase. Table 3.1 indicates possible natural attenuation 

and mobilization pathways for nickel. Details of the types 

of analytical measurements that may be conducted on 

sampled ground water and aquifer sediments to assist in 

identifying the attenuation mechanism(s) are discussed in 

the following paragraphs.

Aqueous Measurements

The total concentration of nickel in aqueous samples can 

be determined by an array of methods ranging significantly 

in sensitivity, detection limits, and accuracy. For aqueous 

systems, nickel can be measured by flame/graphite furnace atomic absorption (FAAS and GFAAS, respectively), 

inductively coupled plasma atomic emission spectrometry 

(ICP-AES) or mass spectrometry (ICP-MS), colorimetry, 

ion chromatography, and electrochemical methods (Stoeppler, 1980). The standard colorimetric method for nickel 

is the dimethylglyoxime (DMG) method (Amacher, 1996). 

Ion chromatography works well for nickel in determining 

total nickel in soil digestion solutions (Basta and Tabatabai, 

1990). Electrochemical methods (e.g., anodic stripping 

voltammetry, platinum electrode differential oscillopolarography, or differential pulse polarography) are well suited for 

aqueous samples and often employ DMG-coated electrodes 

to concentrate nickel for better sensitivity (Stoeppler, 1980). 

Of the list above, FAAS, GFAAS, ICP-OES, or ICP-MS are 

the most common methods employed. 

For ground water with elevated concentrations of 

dissolved organic carbon (e.g., landfill leachates) or 

known organic co-contaminants such as EDTA, it may 

be necessary to determine the chemical speciation of 

dissolved nickel. Geochemical speciation models (e.g., 

MINTEQA2, PHREEQC, EQ3/6) may be employed to 

assist in determining aqueous nickel speciation, but the 

Table 3.1 Natural attenuation and mobilization pathways for nickel.

Attenuation Processes Mobilization Processes Characterization Approach

Precipitation of Ni as carbonate, 

layered double hydroxide, or 

phyllosilicate phase in oxidized/

reduced systems; precipitation of 

Ni as a sulfide in sulfate-reducing 

systems

Dissolution of Ni precipitates 

due to decreased pH; dissolution of NiS due to shift 

from reducing to oxidizing 

conditions.

Evaluation of dissolved Ni concentration in ground 

water. Determination of total Ni in the solid matrix 

and suspected components in Ni-bearing precipitate. 

Evaluation of mineral solubility relative to ground-water chemistry and published solubility constants.

Co-precipitation of Ni as a trace 

component in oxyhydroxides or 

sulfides of iron or manganese

Dissolution of host oxyhydroxide due to decrease in pH or 

shift from oxidizing to reducing 

conditions; dissolution of host 

sulfide due to shift from reducing to oxidizing conditions.

Evaluation of Ni concentration in ground water and in 

solid matrix. Evaluation of host precipitate formation 

relative to existing ground-water chemistry; determination of host mineral content in aquifer sediments 

via mineralogical characterization. Evaluation of Ni 

solid-phase partitioning using sequential extraction 

methodologies.

Adsorption of Ni to iron oxyhydroxides, iron sulfides, or other 

mineral surfaces 

Desorption due to low pH, high 

competing cation concentrations, or high DOC concentrations for oxyhydroxides and 

sulfides. Reductive dissolution 

of iron hydroxides or oxidative 

dissolution of iron sulfides.

Evaluation of Ni concentration in ground water and in 

solid matrix. Evaluation of Ni solid-phase partitioning 

using sequential extraction methodologies. Batch 

and column testing to determine Ni uptake behavior 

and capacity of site-specific aquifer materials under 

variable geochemical conditions.
26

accuracy of existing codes and/or associated geochemical 

databases for assessing complexation with dissolved 

organic carbon has been questioned (e.g., Christensen 

and Christensen, 2000). These computer speciation codes 

also require, at a minimum, the concentrations of major 

anions, major cations, total organic carbon (or specific 

species of organic compounds), temperature, and pH. 

Direct determination of the fraction of organic-complexed 

nickel may be accomplished through analytical fractionation 

of nickel using various exchange resins (e.g., Christensen 

and Christensen, 2000; Jian and Presley, 2002). As noted 

by Jiann and Presley (2002), the approach to sample 

preservation prior to separation may depend on whether 

fractionation can be conducted in the field or at a later time 

in the laboratory. Voltammetric measurements may also be 

employed to examine the relative distribution and stability 

of nickel complexes with dissolved organic compounds in 

water (e.g., Van den Berg and Nimmo, 1987; Bedsworth and 

Sedlak, 1999). Supporting data for these measurements 

would include determinations of total dissolved organic 

carbon along with specific organic constituents suspected 

in the ground-water plume. In addition, while regulatory 

requirements stipulate that unfiltered ground-water 

samples be analyzed to support regulatory decisions at 

a contaminated site, it may be necessary to also collect 

filtered samples to help define the process(es) controlling 

contaminant mobility. The use of 0.45 μm pore size filters 

is common as an arbitrary cutoff point to differentiate 

between dissolved and particulate phases in water 

samples. However, caution is recommended when using 

this approach, particularly for Fe and Al and other elements 

that may be associated with Fe or Al particles (including Ni) 

that could pass through 0.45 μm filters. The use of filters 

with pore sizes less than 0.1 μm will generally provide a 

better assessment of the dissolved vs. particulate load in 

ground water.

It has also been observed that nickel may be leached from 

certain grades of stainless steel well casing/screen materials under chemical conditions that may be encountered in 

contaminant plumes. Two published studies have provided 

detailed evaluation of the extent of nickel (and chromium) 

leaching that may occur for type 304 or 316 stainless steel 

screens (Hewitt, 1994; Oakley and Korte, 1996). Oakley 

and Korte (1996) provide a site-specific example of how 

elevated nickel concentrations derived from continuous 

leaching of well screen materials may be falsely identified 

as a component of a ground-water plume. This suggests 

that careful consideration should be given to the types of 

well screen materials and sampling protocols employed 

relative to the assessment of potential contaminants of 

concern within a plume.

Solid Phase Measurements

The implementation of an analytical approach to identify 

nickel speciation in aquifer sediments is a challenging process. The accuracy of the analytical finding is dependent on 

the method of sample collection/preservation and the tools 

used to identify the mechanism of nickel partitioning. It is 

recommended that the analytical protocol be designed to 

address the potential redox sensitivity of the solid phase(s) 

to which nickel may be partitioned (e.g., sulfides in reduced 

sediments). Tools to evaluate the mechanism of nickel 

solid phase partitioning range in complexity from relatively 

simple chemical extractions to advanced spectroscopic 

techniques.

Bulk solid phase nickel concentration can be determined 

directly on the solid sample by X-ray fluorescence spectrometry, neutron activation analysis, or following chemical 

digestion and analysis of nickel in the resultant liquid phase. 

A variety of digestion or extraction methods can be found in 

the literature (Amacher, 1996). Neutron activation analysis 

is not commonly employed due to the scarcity of neutron 

sources required to irradiate the sample. X-ray fluorescence 

is the most attractive approach due to the relative ease 

of sample preparation, which may be conducted with the 

sample in its original state or following fusion with lithium 

metaborate. When combined with the determination of 

other major or trace elements in the solid sample, this 

provides an initial step for assessing possible association 

of nickel with various solid phase components. This type 

of analysis can be conducted on the bulk sample as well 

as at a microscopic level using wavelength (electron microprobe) or energy dispersive spectroscopy coupled to a 

scanning or transmission electron microscope. Microscopic 

examination allows one to better differentiate whether nickel 

may be distributed across a number of different mineral 

phases within the solid sample or primarily associated with 

a discrete phase. There are limitations to this approach 

(Pye, 2004), a significant one being that the analysis does 

not necessarily provide unique mineral identification necessitating the collection of supporting mineralogical and 

chemical data.

More detailed information on the specific partitioning 

mechanism(s) controlling nickel solid phase speciation is 

typically required to adequately support site assessment 

for potential reliance on natural attenuation as part of a site 

remedy. There have been many applications of sequential 

extraction schemes to assess the speciation of solid phase 

nickel (e.g., Tessier et al., 1979; Ryan et al., 2002; Peltier et 

al., 2005; Buanuam et al., 2006). As discussed in the cited 

reports, sequential extraction methods provide a useful tool 

to assist in determining the chemical speciation of trace 

metals in soils/sediments, but essentially all documented 

methods show analytical limitations in selectively extracting nickel and other metals associated with specific solid 

components. Where feasible, it is recommended that complimentary analytical techniques be employed to confirm the 

accuracy of nickel speciation (e.g., D’Amore et al., 2005; 

Manceau et al., 2006) or the accuracy of the extraction of 

a targeted phase(es) for a given extractant (e.g., Shannon 

and White, 1991; Ryan et al., 2002; Peltier et al., 2005). As 

an example, Peltier et al. (2005) have demonstrated that 

a common extraction method employed to target metals 

associated with easily reducible iron (hydr)oxides may also 

dissolve iron sulfides that may be present. The results from 

this analysis may lead to misidentification of a nickel association with iron (hydr)oxides, resulting in the development of 

a conceptual site model that misrepresents the site-specific 
27

attenuation process. Under reducing conditions, it is also 

critical that aquifer sediments be sampled and processed 

in a manner that prevents exposure to oxygen prior to extraction in order to limit oxidation of reduced minerals (e.g., 

iron sulfides) that may host nickel. Determination of the 

host mineral phase(es) dissolved for each extraction step 

is recommended, along with the use of surrogate Ni-bearing phases spiked into the sediment to confirm accuracy 

of the procedure (e.g., Rudd et al., 1988). The choice of 

appropriate nickel surrogate phases would be governed by 

site-specific geochemical conditions or characterization of 

the mineralogy of the aquifer sediment.

Long-term Stability and Capacity

The stability of attenuated nickel will largely depend on the 

stability of site-specific geochemical conditions through 

time. For example, if nickel attenuation follows a pathway 

of coprecipitation with iron sulfide, then the long-term 

stability of attenuated nickel will depend, in part, on the 

persistence of reducing conditions. If ground-water redox 

conditions were to shift to oxidizing conditions, nickel might 

be expected to release from the solid phase. It is therefore 

important to understand the attenuation mechanism(s) so 

that geochemical triggers for mobilization can be anticipated 

and incorporated into evaluations of long-term monitoring 

data. For any proposed and identified attenuation mechanism, there will exist possible scenarios whereby remobilization could occur (i.e., changes in pH or Eh). It will be 

essential to explore the likelihood of such changes in site 

geochemistry and the sensitivity of the attenuation pathway 

to changes in the prevailing geochemical conditions.

Quantifying the attenuation capacity (as defined in Volume 

1) will also require an understanding of the specific attenuation pathway(s). Attenuation capacity, for example, could 

be related to the extent that pH is buffered, the availability 

of sorptive sites in aquifer materials, or to the supply of 

electron donors needed to sustain microbially mediated redox conditions. For any proposed attenuation mechanism, 

there will be assumptions built into capacity estimations, 

so it is recommended that uncertainty analysis accompany 

capacity calculations.

Tiered Analysis

Determination of the viability of nickel remediation in ground 

water via monitored natural attenuation will depend upon 

proper assessment of contaminant loading to the aquifer 

and prevailing geochemistry and mineralogy within the 

contaminant plume and the down gradient zone prior to 

the point(s) of compliance. The goal of site assessment is 

to demonstrate the process(es) controlling nickel sequestration onto aquifer solids and the long-term stability of 

solid phase nickel as a function of existing and anticipated 

ground-water chemistry. The following tiered analysis 

structure for site characterization provides an approach to 

evaluate candidate sites and define the potential limitations 

of MNA as part of a remedy for ground-water cleanup.

Tier I - Site characterization under Tier I will involve demonstration that the plume is static or shrinking, has not reached 

compliance boundaries, and does not impact existing water 

supplies. Once this is established through ground-water 

characterization, evidence is collected to demonstrate Ni 

partitioning to aquifer solids within the plume. If natural 

attenuation processes are active throughout the plume, 

then there should be an observed increase in solid phase 

concentrations within regions of the plume with higher 

aqueous concentrations, e.g., near the source term. This 

field partitioning data may be supplemented by geochemical 

modeling that incorporates measured water chemistry (e.g., 

pH, Eh, and major ion chemistry) throughout the plume to 

assess the potential for solubility control by a nickel precipitate such as phyllosilicate or sulfide phase. Identification of active sequestration to prevent nickel migration in 

ground-water provides justification for proceeding to Tier 

II characterization efforts.

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

of attenuation are determined. Estimates of a site attenuation rate(s) can be assessed via a well transect along 

the ground-water flow path. In addition, time-series data 

may be collected at one or more monitoring points within 

the plume. This information will allow assessment of the 

relative timescales for contaminant immobilization and fluid 

transport and determination of whether remediation objectives can be met within the required regulatory time frame. 

In addition, the mechanism(s) for attenuation need to be 

identified under this stage of site characterization. This 

effort may require determination of the chemical speciation 

of aqueous and solid phase Ni, which may be approached 

according to the following scheme:

1) Determination of nickel solution speciation via direct analytical measurements in combination with 

speciation calculations based on characterized 

ground-water chemistry;

2) Calculation of the saturation state of ground water 

relative to measured aqueous chemistry complimented by the possible isolation of discrete Ni 

mineral phases via density separations (or other 

schemes) in regions of the aquifer with highest 

solid phase concentrations;

3) Determination of aquifer mineralogy to determine 

the relative abundance of components with documented capacity for Ni sorption (e.g., Amonette, 

2002);

4) Identification of nickel association(s) with the 

various solid phase components of aquifer solids 

through combination of chemical extractions with 

microscopic/spectroscopic confirmation of phase 

associations, and;

5) Demonstration of concurrence between the site 

conceptual model and mathematical model(s) that 

describe nickel removal mechanism(s).

It is recommended that identification of nickel chemical speciation in aqueous and solid matrices be conducted using 

samples collected in a manner that preserves the in-situ 

distribution of dissolved nickel and mineralogy and prevents 

loss of nickel from aqueous samples (e.g., due to oxidation 

and precipitation of ferrous iron in anoxic ground water). 
28

The demonstration of concurrence between conceptual and 

mathematical models describing nickel transport will entail 

development of site-specific parameterization of the chemical processes controlling nickel solid phase partitioning.

Tier III - Once the partitioning mechanism(s) have been 

identified for the site, the subsequent characterization effort under Tier III will involve determination of the stability 

of immobilized Ni and the capacity of the aquifer to sustain 

continued uptake. It is recommended that the stability of 

immobilized Ni be tested based on the anticipated evolution 

of ground-water chemistry concurrent with plume shrinkage. 

For example, changes in ground-water pH can exert a significant influence on Ni adsorption or precipitate solubility. 

Therefore, it is recommended that sediment leach tests be 

conducted to characterize the magnitude of Ni mobilization 

as a function of pH for a ground-water chemistry representative of site conditions. It is recommended that the capacity 

for Ni uptake onto aquifer solids be determined relative to 

the specific mechanism(s) identified in Tier II. For example, 

if site characterization under Tier II indicated that precipitation of Ni sulfide due to microbial degradation of organic 

compounds coupled with sulfate reduction occurs within the 

aquifer, then it is recommended that the mass distribution 

of organic carbon and sulfate to support this reaction within 

the aquifer be determined. This site-specific capacity can 

then be compared to Ni mass loading within the plume in 

order to assess the longevity of the natural attenuation 

process. If site-specific tests demonstrate the stability of 

immobilized Ni and sufficient capacity within the aquifer to 

sustain Ni attenuation, then the site characterization effort 

can progress to Tier IV. For cases where contaminant stability is sufficient but aquifer capacity is insufficient for capture 

of the entire plume, then a determination of the benefits of 

contaminant source reduction may be necessary.

Tier IV – Finally, under Tier IV a monitoring plan is established along with contingency plans in the event of MNA 

failure. It is recommended that the monitoring plan be 

designed to establish both continued plume stability and to 

identify changes in ground-water chemistry that may lead 

to re-mobilization of attenuated Ni. The specific chemical 

parameters to be monitored will include those identified 

under Tier III that may halt Ni partitioning to aquifer sediments and/or result in solubilization of either discrete Ni precipitates or aquifer minerals that sequester Ni from ground 

water. For example, solution phase parameters that could 

alter either Ni precipitation or adsorption include increases 

in soluble organic carbon in combination with changes in 

ground-water pH. In contrast, the concentration of dissolved 

iron or sulfate may indicate the dissolution of an important 

sorptive phase within the aquifer (e.g., reductive dissolution 

of iron oxides or oxidative dissolution of sulfides). Changes 

in these parameters may occur prior to observed changes 

in solution Ni and, thus, serve as monitoring triggers for 

potential MNA failure. In this instance, a contingency plan 

can be implemented that incorporates strategies to arrest 

possible plume expansion beyond compliance boundaries. 

Possible strategies to prevent plume expansion include 

pump and treat operations, installation of reactive barriers 

to enhance uptake capacity perpendicular to the direction 

of plume advance, or enhancement of natural attenuation 

processes within the aquifer through the injection of soluble 

reactive components.

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33

Copper

Richard T. Wilkin

Occurrence and Distribution

Copper is a minor element in the earth’s crust, ranking 

25th in abundance and having an average concentration of 

25 mg kg-1 (Wedepohl, 1995). Crustal copper concentrations vary as a function of lithologic type and proximity to 

hydrothermal deposits of copper and other base metals. 

Mafic and ultramafic rocks such as basalts are usually 

more enriched in copper compared to rocks that make up 

continental crust such as granites and rhyolites. Median 

concentrations of copper reported in a survey of sediments 

and soils range from 7 to 35 mg kg-1 (Reimann and Caritat, 

1998). Ores of copper are highly diverse and range from: 

native copper deposits hosted in sulfur-poor basaltic and 

andesitic rocks; copper sulfides hosted in layered mafic 

intrusions, altered porphyritic rocks, and submarine massive 

sulfide bodies; and, copper oxides, carbonates, and sulfates 

formed in supergene deposits through the weathering of 

primary sulfide deposits (Guilbert and Park, 1986).

It is believed that the ancient civilizations of Mesopotamia 

(region of modern Iraq) made use of native copper and gold, 

possibly as far back as ca. 8700 BC. Indeed the exploitation 

of natural deposits of copper and gold, which are soft metals that can be hammered into shape without heat, marked 

the transition from the Stone Age to more modern ways of 

life (Diamond, 1997). Typical copper-bearing minerals include native copper, sulfides (chalcopyrite, CuFeS2; bornite, 

Cu5FeS4; chalcocite, Cu2S; covellite, CuS; digenite, Cu9S5), 

sulfosalts (tetrahedrite, Cu12Sb4S13), carbonates (malachite, 

Cu2(OH)2CO3; azurite, Cu3(OH)2(CO3)2), and oxides (tenorite, CuO). In the US, the principal copper mining states are 

Arizona, Utah, and New Mexico. In 2004, domestic mine 

production of copper was about 1.16 million tons, or about 

8% of world mine production. While the US is the world’s 

second-largest producer of copper, it is the world’s largest 

copper-consumer. The principal modern use of copper is 

as an electrical conductor. Alloys of copper (e.g., brass 

and bronze) are used in jewelry, sculptures and for minting coins. Copper also has broad uses as an agricultural 

poison and as an algaecide in water purification.

Plume Characteristics

Copper has five possible oxidation states (0, +1, +2, +3, 

and +4). Under most conditions, copper is present in 

aqueous solution as the divalent cation, Cu2+, or as Cu(II) 

hydroxide or carbonate complexes. However, copper is 

not especially mobile in aquatic environments due to the 

relatively low solubility of Cu(II)-bearing solids and high 

affinity of copper for mineral and organic surfaces. Certain organic compounds are able to keep copper soluble 

by ligand complexation. Hence, copper does not typically 

enter ground water except under conditions of low pH or 

high ligand concentrations. Where present, copper contamination in soils and ground water stems primarily from 

mining activities, metal production, wood production, fertilizer production, and combustion of fossil fuels and wastes 

(e.g., Bocheńska et al., 2000; Zagury et al., 2003).

Copper is not included on the CERCLA Priority List of 

Hazardous Substances, which is based on the frequency 

of occurrence of specific contaminants at National Priorities 

List (NPL) sites and their potential threat to human health. 

An internet search showed, however, that in 2005 copper 

was listed as a potential contaminant of concern (COC) in 

ground water at 287 NPL sites in EPA Regions 1-10.

Remedial Technologies 

The primary techniques for dealing with copper-contaminated soils involve immobilization and/or extraction. Immobilization involves binding copper or other heavy metals to 

the soil matrix by solidification or stabilization. In this way, 

contaminated soils become less soluble, and hazardous 

compounds are prevented from entering ground water or 

surface water. Extraction involves a combination of processes to actually remove heavy metals from soil, for example, 

soil washing whereby metals are transferred into solution 

via solubilization by acids, bases, or chelating agents. In 

contrast to soils remediation, there are comparatively few 

examples of ground-water remediation demonstrations 

that focus on copper. In situ bioremediation to promote 

bacterial sulfate-reduction and consequent precipitation 

of insoluble copper sulfides has been proposed (Dvorak et 

al., 1992; Steed et al., 2000; Tabak et al., 2003). Permeable reactive barriers that are designed to intercept and 

treat contaminated ground water could be appropriate for 

dealing with copper contamination. Woinarski et al. (2003) 

discuss the application of a natural zeolite (clinoptilolite) in 

reactive barriers for removing copper via ion exchange (see 

also Inglezakis et al., 2003 and Park et al., 2002). Other 

reactive media explored in laboratory studies for treating 

copper include zerovalent iron (Wilkin and McNeil, 2003) 

and municipal compost (Waybrant et al., 1998).
34

Regulatory Aspects

The USEPA has set the Maximum Concentration Limit Goal 

(MCLG) of copper in drinking water at 1.3 mg L-1 (USEPA, 

2006a; http://www.epa.gov/waterscience/criteria/drinking/

dwstandards.pdf). Copper is a trace element essential for 

good human health. It is part of the prosthetic groups of 

many proteins and enzymes and thus is essential to their 

proper function. Potential health effects from ingesting 

water with high concentrations of copper include gastrointestinal distress, and potential damage to the liver and 

kidneys. For non-potable water sources, ambient water 

quality criteria (AWQC) that are protective of aquatic life 

may serve as alternative cleanup goals. For copper, current 

statutes list both acute and chronic criteria for fresh waters 

as 0.016 mg L-1 and 0.011 mg L-1, respectively, for a water 

hardness of 100 mg L-1 (USEPA, 2006b; http://www.epa.

gov/waterscience/criteria/nrwqc-2006.pdf). Adjustments 

to these criteria are to be applied for waters with different 

hardness. An example of where this criterion may apply 

is a site where contaminated ground water discharges to 

surface water.

Geochemistry and Attenuation Processes

Aqueous Speciation

Copper complexes are possible in the +1, +2, +3, and +4 

valence states. Copper(III) and (IV) complexes are rare 

and unstable in water. Cu(I) complexes are present under 

reducing conditions but in general cuprous ions are highly 

insoluble in water. Cu(II) is the main oxidation state for 

soluble complexes of copper in aquatic environments. Cu(II) 

forms complexes with both hard (e.g., CO3

2-, SO42-, OH-

, 

and Cl-) and soft (e.g., S2-, I-) bases (Stumm and Morgan, 

1996). The stereochemistry of Cu(II) principally involves 

distorted tetragonal (coordination number 4) or octahedral 

(coordination number 6) configurations. Cu(II) complexes 

are subject to the Jahn-Teller effect that acts to stabilize 

species with the d 9 electronic configuration that are present 

in tetragonal or octahedral coordination. 

In pure water, Cu2+ is the predominant ion below pH 7. 

Above this pH, the species CuOH+, Cu(OH)2

0, and CuO22-

become increasingly important. Because of uncertainties 

in the estimates and measurements of the thermodynamic 

constants for copper complex formation, the speciation 

of copper in natural waters is not known in detail (e.g., 

Boyle, 1979; Leckie and Davis,1979; Baes and Mesmer, 

1976). Most models, however, predict that Cu2+ is a small 

fraction of the total copper concentration in freshwater and 

seawater systems and that complexed forms of copper are 

dominant.

A survey of references that report hydrolysis constants for 

Cu2+ shows considerable variability in species identified and 

in their formation constants (see e.g., Baes and Mesmer, 

1976; Leckie and Davis, 1979 and references therein). 

Figure 4.1 shows the pH-dependent distribution of Cu2+ hydrolysis species based on the Lawrence Livermore National 

Laboratory thermodynamic database (thermo.com.v8.r6+), 

along with the pH-dependent solubility of tenorite (CuO) and 

the metastable solid, Cu(OH)2, based on data in Hidmi and 

Edwards (1999). Both CuO and Cu(OH)2 are insoluble at 

neutral to alkaline pH. Below pH of 6 to 7, dissolution of 

these phases would yield Cu2+ concentrations above the 

MCL of 1.3 mg L-1 (10-4.7 molal). Stable complexes of Cu2+

with SO4

2- (CuSO40) and CO32- (CuCO30) may contribute 

significantly to total copper at anion concentrations typically encountered in ground water. In environments with 

high ammonia concentrations, copper can be significantly 

bound to ammonia at pH>6.

Copper may be strongly complexed by dissolved organic 

matter (e.g., Smolyakov et al., 2004). Low molecular 

weight, dissolved organic complexes are generally highly 

mobile and able to transport copper in aquifer materials 

and soils (Han and Thompson, 2003; Christensen et al., 

1999). Christensen et al. (1999) report that >85% of total 

copper was bound to dissolved organic carbon complexes 

in leachate with comparatively low dissolved organic carbon 

concentrations of <40 mg C L-1.

Figure 4.1 Solubility of copper oxide and copper 

hydroxide as a function of pH in the system 

Cu-O-H at 25 °C.

Solubility

An Eh-pH diagram for copper is shown in Figure 4.2. 

Inspection of this diagram indicates that at the specified 

conditions Cu2+ is the soluble form of copper at pH<6 and 

in moderately to highly oxidizing systems. Consequently, 

upon weathering, copper is likely to be more mobile under 

acidic rather than alkaline conditions (see, e.g., Paulson 

and Balistrieri, 1999). Copper hydroxycarbonate (malachite) has a narrow stability field at near-neutral pH and at 

moderately to highly oxidizing conditions. With increasing 

inorganic carbon concentrations, the malachite stability field 
35

would expand. At neutral to alkaline pH (>7) copper oxides 

are stable. With progressively more reducing conditions, 

cuprous oxide and elemental copper develop broad stability fields. Finally, in highly reducing and sulfidic environments, copper sulfides (chalcocite and covellite) are stable 

over a wide pH range. Aquifer materials usually contain 

some organic matter as well as sulfate from ground water. 

Microbial degradation of organic matter can be coupled to 

sulfate reduction with the production of hydrogen sulfide. 

Because of the extreme insolubility of copper sulfides, no 

complexing ligand can compete with hydrogen sulfide or 

metal sulfide surfaces for copper (Rose, 1989). In addition, there are several fairly common copper-iron-sulfur 

minerals such as chalcopyrite (CuFeS2), bornite (Cu5FeS4), 

and cubanite (CuFe2S3). Note that in solutions with high 

chloride concentrations (>1 M), the field of soluble copper 

in Figure 4.2 expands substantially due to the very strong 

nature of copper chloride complexes.

Figure 4.2 Eh-pH diagram for copper at 25 °C (total 

inorganic carbon = 10-2 molal; total sulfur = 

10-3 molal; total copper = 10-5 molal).

Cavallaro and McBride (1980) found that in alkaline soils 

copper was present as Cu(OH)2 and with progressive 

aging, copper solubility decreased consistent with the formation of Cu2(OH)2CO3. In general, they concluded that 

Cu2+ solubility in soil is highly correlated to pH. Dudley et 

al. (1991) proposed the formation of CuO in soil reacted 

with an extract of acid mine waste. In both the studies of 

Cavallaro and McBride (1980) and Dudley et al. (1991), 

proposed copper phase associations are based on a comparison of observed pH-dependent copper concentrations 

with solubility estimates based on thermodynamic data. 

Indeed, Leckie and Davis (1979) suggest that in most soil 

environments malachite and tenorite are the most important copper-bearing phases, with Cu(OH)2 present as a 

metastable precursor to malachite and tenorite.

The stability relationships between copper hydroxycarbonates and oxides are shown in Figure 4.3 in terms of pH 

and CO2 fugacity. Note that at high pH and CO2 fugacity, 

aqueous copper carbonate complexes predominate over 

hydroxyl complexes. The diagram illustrates that over the 

pH and f CO2 conditions in most ground water systems, 

tenorite and malachite are the expected stable copper 

minerals.

Adsorption

McBride and Bouldin (1984) examined the solid-phase 

properties of copper in copper-contaminated soil. They 

concluded that long-term reaction of copper with calcareous soil failed to convert copper into a form unavailable to 

plants. Chemical extraction tests suggested that copper 

was mainly present in a non-exchangeable form easily 

dissolved by organic chelating compounds. An analysis 

of pH-dependent solubility data revealed that copper was 

present as an inorganic form in the soil, possibly tightly 

adsorbed on surfaces as hydroxyl or hydroxycarbonate species. Greater than 99.5% of the copper in the soil solution 

was complexed, probably with soluble organic compounds 

(McBride and Bouldin, 1984). In contrast, Cavallaro and 

McBride (1978) found that low pH soils are less effective 

in retaining Cu2+ compared to neutral soils and calcareous 

soils. They concluded that this behavior was in part related 

to increased competition at low pH for organic functional 

groups by aluminum and/or protons thus reducing the ability 

of Cu2+ to be adsorbed onto solid organic matter.

Figure 4.3 Solubility and speciation of copper as a 

function of pH and log fugacity of CO2(gas) 

at 25 °C. Solid lines separate stable phases 

and dotted lines separate aqueous species 

(total copper = 10-5 molal). Diagram drawn 

using thermodynamic data from MINTEQA2.
36

Based on the solubility and sorption behavior of copper, it is 

expected that over a wide range of geochemical conditions 

copper will be effectively stable in the solid phase of soils 

and sediment materials. Copper has a strong affinity for 

the surfaces of iron oxides and hydroxides (e.g., Benjamin 

and Leckie, 1981; Robertson and Leckie, 1998; Martínez 

and McBride, 1998), clays (e.g., Pickering, 1980; Farqhar 

et al., 1997; Morton et al., 2001), sulfides (e.g., Pattrick et 

al., 1997; Parkman et al., 1999), and organic matter (e.g., 

Sauvé et al., 1997; Schilling and Cooper, 2004). As well 

as being less soluble, Cu2+ is more strongly adsorbed to 

mineral substrates than Zn2+, Ni2+, and Cd2+.

Benjamin and Leckie (1981) examined the pH-dependent 

sorption of copper, zinc, and lead onto hydrous ferric oxide. 

For these metals and for a range of precipitate loadings, 

the adsorption edge position, the pH at which half the 

metal was sorbed and half the metal remained in solution, 

increased in the order Pb<Cu<Zn<Cd. In this study, the pH 

at which 50% of copper was taken up by the surface of the 

iron precipitates ranged from about 5.0 to 5.5. At pH>6, 

copper was essentially completely removed from solution. 

Khaodhiar et al. (2000) observed nearly identical pH-dependent behavior for copper adsorption onto iron oxide coated 

sand grains. Martínez and Motto (2000) determined the 

pH at which metal amended soils began to release copper via an acid titration method. Interestingly, they found 

that copper was released at about pH 5.5 ± 0.2 which is in 

good agreement with the adsorption studies, and further 

reinforces the notion of reversible sorption processes and 

potential copper mobility at low pH.

Redox Chemistry

Equilibrium between cupric and cuprous ions can be represented by the equation:

Cu2+ + e- = Cu+ (log K = 2.72)

In natural systems the stable solid in very reducing conditions is expected to be cuprous sulfide (Cu2S, chalcocite, 

see Figure 4.2). As the Eh increases there is a narrow 

window in which cupric sulfide (CuS, covellite) becomes 

important. Further increases in Eh can lead to the formation of elemental copper. So in general the solubility and 

speciation of copper are determined by redox equilibria 

of sulfur and copper and the strength of available ligands. 

Experimental studies of Cu(I) complexation by chloride and 

bisulfide are presented in Xiao et al. (1998), Thompson and 

Helz (1994), Mosselmans et al. (1999), and Mountain and 

Seward (1999, 2003). Luther et al. (2002) show that the 

reduction of Cu(II) to Cu(I) occurs in sulfidic solutions prior 

to the precipitation of copper sulfides.

Colloidal Transport 

Recent studies are consistent in demonstrating that copper 

in ground water is frequently associated with colloids that 

appear to be organic in nature (Sañudo-Wilhelmy et al., 

2002; Jensen et al., 1999; Freedman et al., 1996; Pauwels 

et al., 2002). The association between metals and ground 

water colloids is evident both in uncontaminated groundwater (e.g., Sañudo-Wilhelmy et al., 2002) and in contaminated landfill leachates and in ground water impacted 

by mining districts (e.g., Jensen et al., 1999; Pauwels et 

al., 2002). Jensen et al. (1999) found that 86-95% of total 

copper in landfill leachate was associated with small-size 

colloidal matter and organic molecules. They concluded 

that most metals, including copper, present in the colloidal 

forms would have been sampled in the dissolved fraction 

if the commonly employed filter size of 0.45 μm had been 

used, since only negligible amounts of metal were found 

with colloids >0.40 μm. Pauwels et al. (2002) found that 

the mobility of copper in ground water impacted by the 

oxidative dissolution of massive sulfide deposits in the 

Iberian Pyrite Belt (Spain) was especially enhanced due 

to complexation with organic matter and/or adsorption onto 

colloids. For example, measured concentrations of copper 

were 108 to 109 times greater than concentrations modeled 

assuming equilibrium with respect to sulfide minerals (e.g., 

chalcopyrite).

Site Characterization

Copper mobility in ground water and the risk of copper 

exposure to plants, animals, and/or humans is governed by 

the total concentration of copper, the distribution of copper 

species in water, and the nature of copper partitioning in the 

solid phase. The development of site-conceptual models 

for predicting the long-term fate of copper at a contaminated site will require information on the distribution and 

concentration of copper in the aqueous phase and the solid 

phase. Table 4.1 indicates possible natural attenuation and 

mobilization pathways for copper.

Quantitative measurement of copper concentrations in 

aqueous solutions is typically carried out using inductively 

coupled plasma optical emission spectroscopy (ICP-OES), 

inductively coupled plasma mass spectroscopy (ICP-MS), 

or atomic absorption spectroscopy (AAS). Some of the 

unique features of determining copper concentrations in 

natural waters are discussed in Boyle (1980) and SañudoWilhelmy et al. (2002). Input data to geochemical codes 

(e.g., MINTEQA2, PHREEQC, EQ3/6) for determining 

aqueous speciation also require, at a minimum, the concentrations of major anions, major cations, total organic 

carbon, temperature, and pH. The total concentration of 

copper in soils, sediments, and aquifer materials may be 

determined by X-ray fluorescence (XRF) spectroscopy, or 

by chemical analysis after digestion in mineral acids.

While regulatory requirements stipulate that unfiltered 

ground-water samples be analyzed to support regulatory decisions at a contaminated site, it may be necessary to also collect filtered samples to help interpret that 

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 Cu) that could pass through 0.45 μm 

filter papers. The use of filter papers with pore sizes less 

than 0.1 μm will generally provide a better assessment of 

the dissolved vs. particulate load of a ground water or a 

surface water sample.
37

Hickey and Kittrick (1984) examined the chemical partitioning of copper in soils and sediments containing high levels 

of heavy metals using the selective extraction approach 

developed by Tessier et al. (1979). In this study, copper 

was assigned to five operationally defined geochemical 

fractions: exchangeable, bound to carbonates, bound to 

Fe- and Mn-oxides, bound to organic matter, and residual. 

This study concluded that copper was the metal most 

significantly associated with organic matter. Compared to 

other heavy metals considered in this study (nickel, and 

zinc), copper displayed a low potential for mobility and 

metal bioavailability (Hickey and Kittrick, 1984).

Long-Term Stability and Capacity

The stability of attenuated copper will largely depend on the 

fluctuation of site-specific geochemical conditions through 

time. For example, if copper attenuation follows a copper 

hydroxycarbonate precipitation pathway, then long-term 

stability of attenuated copper will depend, in part, on the 

persistence of pH conditions. If pH conditions were to shift 

significantly to more acidic values, copper might be expected to release from the solid phase. It is therefore important 

to understand the attenuation mechanism(s) so that geochemical triggers for remobilization can be anticipated and 

incorporated into evaluations of long-term monitoring data. 

For any proposed and identified attenuation mechanism, 

there will exist possible scenarios whereby remobilization 

can occur (i.e., changes in pH or Eh). It will be essential to 

explore the likelihood of such changes in site geochemistry 

and the sensitivity of the attenuation pathway to changes 

in the prevailing geochemical conditions.

Quantifying the attenuation capacity (as defined in Volume 

1) will also necessitate an understanding of the specific 

attenuation pathway(s). Attenuation capacity, for example, 

could be related to the extent that pH is buffered, the 

availability of sorptive sites in aquifer materials, or to the 

supply of electron donors needed to sustain microbially 

mediated redox conditions. For any proposed attenuation 

mechanism, there will be assumptions built into capacity 

estimations, so that uncertainty analysis is recommended 

to support capacity calculations.

Tiered Analysis

Determination of the viability of copper remediation in 

ground water via monitored natural attenuation will depend 

upon proper assessment of contaminant loading to the 

aquifer and prevailing geochemistry and mineralogy within 

the contaminant plume and the down gradient zone prior 

to the point(s) of compliance. MNA may not be appropriate as a site remedy for copper contamination in acidic 

pH, highly oxidizing, and/or DOC-rich environments. The 

goal of site assessment is to demonstrate the process(es) 

controlling copper sequestration onto aquifer solids and 

the long-term stability of solid phase copper as a function 

of existing and anticipated ground-water chemistry. The 

following tiered analysis structure for site characterization 

provides an approach to evaluate candidate sites and define the potential limitations of MNA as part of a remedy 

for ground-water cleanup.

Tier I - Site characterization under Tier I will involve demonstration that the plume is static or shrinking, has not 

reached compliance boundaries, and does not impact 

existing water supplies. Once this is established through 

ground-water characterization, evidence is collected to 

demonstrate Cu partitioning to aquifer solids within the 

plume. If natural attenuation processes are active throughout the plume, then there should be an observed increase 

in solid phase concentrations within regions of the plume 

with higher aqueous concentrations, e.g., near the source 

term. This field partitioning data may be supplemented 

by geochemical modeling that incorporates measured 

water chemistry (e.g., pH, Eh, and major ion chemistry) 

throughout the plume to assess the potential for solubility 

control by copper hydroxide, sulfate, carbonate, phosphate, 

Table 4.1 Natural attenuation and mobilization pathways for copper.

Attenuation Processes Mobilization Processes Characterization Approach

Precipitation of insoluble hydroxycarbonates, carbonates, 

sulfides, and phosphates. In 

general, pH>6 will drive precipitation reactions resulting in 

Cu concentrations below the 

MCL.

Dissolution of carbonates, hydroxycarbonates, and oxides at low pH; 

oxidative dissolution of sulfides 

at low pH and high Eh; complexation/stabilization in the presence 

of DOC.

Evaluation of copper speciation in the aqueous phase. Determination of total Cu in the 

solid matrix. Evaluation of solid phase partitioning using sequential extraction methodologies. Evaluation of long-term sorption capacity/stability.

Sorption to iron hydroxides, 

organic matter, carbonates, 

and sulfides. 

Desorption at low pH; complexation/stabilization in the presence 

of DOC. Reductive dissolution of 

iron hydroxides.

Evaluation of copper speciation in the aqueous phase. Determination of total Cu in the 

solid matrix. Evaluation of solid phase partitioning using sequential extraction methodologies. Batch and column testing to determine 

Cu uptake capacity of site-specific aquifer materials with variable geochemical conditions.
38

or sulfide. This provides justification for proceeding to Tier 

II characterization efforts. 

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

of attenuation are determined. Estimates of a site attenuation rate(s) can be assessed via a well transect along 

the ground-water flow path. In addition, time-series data 

may be collected at one or more monitoring points within 

the plume. This information will allow assessment of the 

relative timescales for contaminant immobilization and fluid 

transport and determination of whether remediation objectives can be met within the required regulatory time frame. 

In addition, the mechanism(s) for attenuation need to be 

identified under this stage of site characterization. This 

effort may require determination of the chemical speciation 

of aqueous and solid phase Cu, which may be approached 

according to the following scheme:

1) Determination of solution speciation via direct analytical measurements (e.g., Martinez et al., 2001; 

Sañudo-Wilhelmy et al., 2002) to aid differentiation 

of uncomplexed (i.e., Cu2+) and complexed (e.g., 

CuCl+, Cu-organic ligand complexes) forms of mobile Cu in combination with speciation calculations 

based on characterized ground-water chemistry;

2) Calculation of the saturation state of ground water 

relative to measured aqueous chemistry complimented by the possible isolation of discrete Cu 

mineral phases via density separations (or other 

schemes) in regions of the aquifer with highest 

solid phase concentrations;

3) Determination of aquifer mineralogy to determine 

the relative abundance of components with documented capacity for Cu sorption (e.g., Amonette, 

2002; Burton et al., 2005); and 

4) Determination of Cu-sediment associations via 

chemical extractions designed to target specific 

components within the aquifer sediment (e.g., Lee 

et al., 2005).

This compilation of information will facilitate identification 

of the reaction(s) leading to Cu immobilization within the 

plume.

Tier III - Once the partitioning mechanism(s) have been 

identified for the site, the subsequent characterization effort 

under Tier III will involve determination of the stability of 

immobilized Cu and the capacity of the aquifer to sustain 

continued uptake. It is recommended that the stability of 

immobilized Cu be tested based on the anticipated evolution 

of ground-water chemistry concurrent with plume shrinkage. 

For example, changes in ground-water pH can exert a significant influence on Cu adsorption or precipitate solubility. 

Therefore, it is recommended that sediment leach tests be 

conducted to characterize the magnitude of Cu mobilization 

as a function of pH for a ground-water chemistry representative of site conditions. It is recommended that the capacity 

for Cu uptake onto aquifer solids be determined relative to 

the specific mechanism(s) identified in Tier II. For example, 

if site characterization under Tier II indicated that precipitation of Cu sulfide due to microbial degradation of organic 

compounds coupled with sulfate reduction occurs within the 

aquifer, then it is recommended that the mass distribution 

of organic carbon and sulfate to support this reaction within 

the aquifer be determined. This site-specific capacity can 

then be compared to Cu mass loading within the plume 

in order to assess the longevity of the natural attenuation 

process. If site-specific tests demonstrate the stability of 

immobilized Cu and sufficient capacity within the aquifer to 

sustain Cu attenuation, then the site characterization effort 

can progress to Tier IV. For cases where contaminant stability is sufficient but aquifer capacity is insufficient for capture 

of the entire plume, then a determination of the benefits of 

contaminant source reduction may be necessary.

Tier IV – Finally, under Tier IV a monitoring plan is established along with contingency plans in the event of MNA 

failure. It is recommended that the monitoring plan be 

designed to establish both continued plume stability and to 

identify changes in ground-water chemistry that may lead 

to re-mobilization of attenuated Cu. The specific chemical 

parameters to be monitored will include those identified 

under Tier III that may halt Cu partitioning to aquifer sediments and/or result in solubilization of either discrete Cu 

precipitates or aquifer minerals that sequester Cu from 

ground water. For example, solution phase parameters 

that could alter either Cu precipitation or adsorption include increases in soluble organic carbon or chloride in 

combination with changes in ground-water pH. In contrast, 

the concentration of dissolved iron or sulfate may indicate 

the dissolution of an important sorptive phase within the 

aquifer (e.g., reductive dissolution of iron oxides or oxidative dissolution of sulfides). Changes in these parameters 

may occur prior to observed changes in solution Cu and, 

thus, serve as monitoring triggers for potential MNA failure. 

In this instance, a contingency plan can be implemented 

that incorporates interventive strategies to arrest possible 

plume expansion beyond compliance boundaries. Possible 

strategies to prevent plume expansion include pump and 

treat operations, installation of reactive barriers to enhance 

uptake capacity perpendicular to the direction of plume 

advance, or enhancement of natural attenuation processes 

within the aquifer through the injection of soluble reactive 

components.

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43

Chromium

Douglas B. Kent, Robert W. Puls, Robert G. Ford

Occurrence and Distribution

 The average crustal abundance of chromium (Cr) is approximately 100 micrograms per gram. Chromium is much 

less abundant than vanadium and manganese (Mn), which 

are its nearest neighbors on the periodic table, but more 

abundant than the other first-row transition elements cobalt 

(Co), nickel, copper, and zinc. Chromium is a group VI 

transition metal, whose name refers to the variety of colors associated with its various oxidation states. Only two 

of these oxidation states are known to occur naturally. In 

minerals that crystallized in the earth’s interior Cr occurs 

exclusively in the plus 3 oxidation state (Cr(III)). In aquatic 

systems at pH values above 2 and in the presence of oxygen, Cr(III) is thermodynamically unstable compared to the 

plus 6 oxidation state (Cr(VI)) and, therefore, is subject to 

oxidation to Cr(VI).

Minerals containing Cr(VI) are rare. The best known is 

the lead (Pb) chromate mineral crocoite (PbCrO4), which 

occurs as spectacular red-orange crystals in the oxidized 

regions of some Pb deposits. It is from this mineral that 

Cr was first isolated and identified. Several other Cr(VI) 

minerals have been identified in evaporite deposits; those 

from the Atacama Desert in South America account for 

most of these minerals. These evaporite minerals include 

chromate (CrO4

2-) or dichromite (Cr2O72-) combined with 

sodium (Na), potassium (K), calcium (Ca), or barium (Ba) 

with varying amounts of sulfate or other anions.

Chromium(VI) has been found to occur naturally in ground 

water underlying the arid Paradise Valley in Arizona, USA, 

at concentrations as high as hundreds to thousands of 

micrograms per liter (μg L-1) (ones to tens of micromoles 

per liter, μM) (Robertson, 1975). The ground water was 

oxic, had alkaline pH values, moderate concentrations of 

dissolved salts, and was likely very old. More recently, 

Cr(VI) has been detected in ground water with similar 

chemical characteristics underlying arid or semi-arid basins 

in California. As discussed further below, oxic ground water 

with neutral-to-alkaline pH values and moderate-to-high 

concentrations of other anions, such as sulfate, possesses 

the chemical conditions that favor the persistence of Cr(VI) 

and promote its mobility. 

Remedial Technologies

The goal of remediation schemes is to reduce the carcinogenic, soluble, and mobile Cr(VI) to the less toxic and less 

mobile Cr(III), which forms minimally soluble precipitates. 

Successful treatment of Cr(VI) hinges upon the formation 

and stability of Cr(III) precipitates. Ex-situ treatment technologies for ground water commonly use pump-and-treat 

approaches with chemical reduction of the Cr(VI) to Cr(III) 

followed by precipitation. In-situ technologies currently 

used for remediation of Cr(VI) contamination employ some 

form of chemical reduction and fixation (e.g., geochemical 

fixation, permeable reactive barriers (PRBs), and reactive 

zones established through chemical injections). There is 

fairly extensive performance data available to evaluate the 

potential for use of PRBs constructed using zerovalent iron 

(Wilkin and Puls, 2003).

Regulatory Aspects

Cr(VI) is considered teratogenic (Abbasi and Soni, 1984), 

mutagenic (Paschin et al., 1983) and carcinogenic (Ono, 

1988). According to the International Agency for Research 

on Cancer (IARC), Cr(VI) is considered a powerful carcinogen and its presence in waters is cause for concern. 

The national primary drinking water standard set by the 

USEPA for total Cr is 0.1 mg L-1 (USEPA, 2006a). The 

state of California maximum contaminant level in water is 

50 μg L-1. No separate drinking water standard for Cr(VI) 

has been established, but a separate standard has been 

discussed for some time. For non-potable water sources, 

ambient water quality criteria (AWQC) that are protective 

of aquatic life may serve as alternative cleanup goals. 

For Cr(VI), current statutes list both acute and chronic 

criteria for fresh waters as 0.016 mg L-1 and 0.011 mg L-1, 

respectively, for a water hardness of 100 mg L-1 (USEPA, 

2006b). Adjustments to these criteria are to be applied for 

waters with different hardness. An example of where this 

criterion may apply is a site where contaminated ground 

water discharges to surface water. 

Geochemistry and Attenuation Processes

Aqueous Speciation

Aqueous speciation of Cr(VI) varies with pH and Cr(VI) 

concentration (Palmer and Puls, 1994). The dominant 

aqueous species above pH 6 is CrO4

2-. The dominant 

species between pH 0 and 6 is HCrO4

-

 at Cr(VI) concentrations below approximately 0.003 moles L-1 (160 mg L-1) 

and Cr2O7

2- at higher Cr(VI) concentrations. The species 

H2CrO4 is dominant only at pH < 0 (i.e., under extremely 

acidic conditions). In the presence of Ca or iron in the plus 
44

3 oxidation state (Fe(III)), solubility studies indicate the 

formation of CaCrO4

0 and FeCrO4-

 as solution complexes 

(Perkins and Palmer, 2000; Baron and Palmer, 1996).

At near-neutral pH values, Cr(III) concentrations in equilibrium with Cr(OH)3 are less than 0.1 μM (approximately 

5 μg L-1); concentrations increase with decreasing pH below 

7 and increasing pH above 11 (Palmer and Puls, 1994). The 

dominant solution species below pH 3.5 is Cr3+; at these low 

pH values Cr(III) can be mobile and appreciable concentrations of Cr(III) may be observed (Seaman et al., 1999). As 

pH increases, the dominant species changes through a sequence of hydrolysis products: CrOH2+, Cr(OH)2

+, Cr(OH)30, 

Cr(OH)4

-

 (Rai et al., 1987). Oligomers like Cr2(OH)2

4+ and 

Cr3(OH)4

5+ may form as metastable species at mildly acidic 

to near-neutral pH values and moderate Cr(III) concentrations (Baes and Mesmer, 1976).

Cr(III) forms strong complexes with a variety of organic 

ligands. Once formed, slow exchange of inner-coordinationsphere ligands around Cr(III) may result in the persistence 

of Cr(III)-organic complexes under conditions where they 

are thermodynamically unstable This may account for 

reports of elevated concentrations of Cr(III) at near-neutral 

pH values and high concentrations of organics in a Cr-contaminated wetland (Mattuck and Nikolaidis, 1996), tannery 

effluent (Walsh and O’Halloran, 1996), and organic soil or 

landfill leachates (Li and Xue, 2001).

Solubility

Chromium(VI) forms soluble compounds with the alkali 

(Na, K) and alkaline earth (Mg, Ca) metal cations that are 

typically present at the highest concentrations in ground 

water. It does form a sparingly soluble salt with barium Ba, 

which is ubiquitous in ground water but usually present at 

trace concentrations; the mineral hashemite (Ba(Cr,S)O4) is 

known from at least one location. The importance of BaCrO4

at sites with Cr(VI)-contamination has been suggested by 

the agreement between ground-water Cr(VI) concentrations 

and those calculated based on BaCrO4 solubility (Baron and 

Palmer, 1996). Equilibrium computations suggest that a 

significant fraction of Cr(VI) could be present as BaCrO4 at 

neutral and alkaline pH values if there is an adequate source 

of Ba (Figure 5.1). Barite (BaSO4) was used as the source 

of Ba used in the computations in Figure 5.1; Ba concentrations in equilibrium with barite are independent of pH in the 

range shown in Figure 5.1. Another potential source of Ba 

in ground water is the dissolution of aluminosilicate minerals. 

These dissolution reactions are typically acid-catalyzed and, 

therefore, the mass of Ba available from these reactions 

should increase with decreasing pH. Chromium(VI) also 

forms sparingly soluble salts with a variety of heavy metal 

ions, of which Pb (PbCrO4) is the most important by virtue 

of the fact that it has been identified at Cr(VI)-contaminated 

sites (Palmer and Wittbrodt, 1991).

Other Cr(VI) compounds can form at very low or very high 

pH values. A laboratory experimental study showed the 

formation of Fe(III) hydroxy chromates mixed with hydrous 

ferric oxide in the pH range 1.5 to 3.5 (Olazabal et al., 1997). 

KFe3(CrO4)2(OH)6, a chromate analogue to jarosite, has 

been identified in acidic, Cr(VI)-contaminated soil (Baron 

et al., 1996). The results of equilibrium calculations, based 

on thermodynamic data for Cr(VI) species and solids from 

Baron and Palmer (1998) suggest that this solid is stable 

below pH 3 (Figure 5.2). Formation of KFe3(CrO4)2(OH)6 in 

the equilibrium computations was driven by high concentrations of Fe(III), which, in turn, were driven by increased 

solubility of soil hydrous ferric oxide with decreasing pH. 

The predicted distribution of Cr(VI) between the aqueous 

and solid phases at low pH is sensitive to the values assumed for the solubility of hydrous ferric oxide and the total 

concentrations of K and Cr(VI).

A Cr(VI) analogue of the Ca, aluminum (Al), carbonate 

mineral ettringite (viz., Ca6[Al(OH)6]

2(CrO4)3•26H2O) was 

observed in Cr(VI)-contaminated concrete, as was a nearly 

pure Cr(VI) hydrocalumite (viz., 3[CaO][Al2O3][CaCrO4]•nH

2O) (Palmer, 2000). Laboratory experiments showed that 

Cr(VI)-ettringite precipitated from Ca- and Al-containing 

solutions at pH values greater than 10 (Perkins and Palmer, 

2000). In systems where there is an excess of carbonate 

over the available Ca, Cr(VI)-ettringite is thermodynamically 

unstable with respect to calcite and gibbsite at high pH. 

Therefore, Cr(VI)-ettringite and related minerals are most 

likely to be important in systems at high pH where there is 

a large excess of Ca over carbonate or where precipitation 

of calcium carbonate is inhibited.

Concentrations of Cr(III) in aqueous systems are limited 

by the low solubility of Cr(OH)3 (Rai et al., 1987). Coprecipitation of Cr(III) with hydrous ferric oxide drives Cr(III) 

concentrations lower and expands the pH range over which 

Cr(III) concentrations are very low (Palmer and Puls, 1994; 

Sass and Rai, 1987).

Figure 5.1 Distribution of 1 mM (millimoles per liter, 

which equals approximately 52 mg L-1) 

chromium(VI) plotted as a function of pH 

in equilibrium with barite. Thermodynamic 

data for aqueous Cr(VI) species from Baron 

and Palmer (1998) and for BaCrO4 and 

BaSO4 from Smith and Martell (1989).
45

Figure 5.2 Calculated distribution of 1mM Cr(VI) (approximately 52 mg L-1) in the presence 

of 1 mM K (approximately 39 mg L-1) and hydrous ferric oxide calculated as a function of 

pH. Thermodynamic data for aqueous and 

solid phase Cr(VI) from Baron and Palmer 

(1998) and for soil hydrous ferric oxide solubility from Lindsay (1988).

Adsorption

Adsorption of Cr(VI) results from chemical reactions between aqueous Cr(VI) species and sites at mineral surfaces. 

Adsorption becomes more favorable with decreasing pH because many of the Cr(VI) adsorption reactions consume H+

and the electrostatic contribution to adsorption onto some 

important adsorbents becomes more favorable at lower pH 

values (Dzombak and Morel, 1990). The intensity of binding of Cr(VI) at the mineral-water interface is intermediate 

between those of strongly binding anions, like phosphate 

and arsenate, and weakly binding anions like sulfate (Davis 

and Kent, 1990). As a result, other anions may out-compete 

Cr(VI) for adsorption sites depending on their concentrations relative to Cr(VI) and the pH. Such competitive effects 

can give rise to complex dependencies of adsorption on 

pH and solute concentrations. This effect is illustrated in 

Figure 5.3. Adsorption of Cr(VI) onto freshly precipitated 

hydrous ferric oxide (HFO) was determined over a range 

of pH values and in the presence of different concentrations of sulfate (Leckie et al., 1984). In the absence of 

sulfate, dissolved concentrations of Cr(VI) decrease with 

decreasing pH and, therefore, so should the mobility of 

Cr(VI). As concentrations of sulfate increase, competition 

for adsorption sites shifts in favor of sulfate and Cr(VI) 

adsorbs less extensively. Thus, Cr(VI) mobility increases 

with increasing concentrations of sulfate and, at constant 

sulfate, it first decreases then increases with decreasing 

pH (Figure 5.3). Other naturally occurring oxyanions, 

such as carbonate and silicate, also compete with Cr(VI) 

for adsorption sites (van Geen et al., 1994; Zachara et al., 

1987). In Cr(VI)-contaminated ground water, competitive 

adsorption with other anions, such as arsenate, phosphate, 

vanadate (Leckie et al., 1984), or organic acids (Mesuere 

and Fish, 1992) may produce similar effects. The influence 

of pH and anion concentration on mobility of Cr(VI) has 

been demonstrated in field-scale transport experiments in a 

sand and gravel aquifer whose adsorption properties were 

dominated by hydrous Fe- and Al-oxide coatings on quartz 

and feldspar grains (Davis et al., 2000; Kent et al., 1995, 

1994). Thus, the mobility of Cr(VI) should be expected to 

vary with aquifer chemistry.

The affinity of soils and sediments for adsorption of Cr(VI) 

varies widely depending on composition. Laboratory column experiments have shown that adsorption of Cr(VI) during transport is greatly enhanced by increasing abundance 

of hydrous ferric oxide in sediments (Martin and Kempton, 

2000; Stollenwerk and Grove, 1985). Adsorption of Cr(VI) 

onto aquifer sediment whose grain surfaces were coated 

with Fe- and Al-containing hydrous oxides was significantly 

less extensive than expected from adsorption onto pure 

hydrous oxides of Fe or Al (Kent et al., 1995). Adsorption of Cr(VI) over a range of pH values onto four different 

soils varied from essentially no adsorption to extensive 

adsorption (Zachara et al., 1989); all four soils had high 

contents of clay-sized minerals and high specific surface 

areas. Results of these studies show that Cr(VI) adsorption onto soils or sediments cannot be predicted a priori

and, therefore, must be determined experimentally using 

site-specific materials.

Figure 5.3 Concentration of dissolved Cr(VI) in equilibrium with Cr(VI) adsorbed on freshly precipitated hydrous ferric oxide. The Fe concentration of 1 mM corresponds to approximately 90 mg FeOOH/L. The 5 μM Cr(VI) 

concentration corresponds to approximately 

260 μg/L. Sulfate concentrations of 1 mM 

and 10 mM correspond to 96 and 960 mg 

SO4/L, respectively Continuous curves calculated from the model for adsorption onto 

hydrous ferric oxide of Dzombak and Morel 

(1990). Experimental data from Leckie et al. 

(1986).
46

Redox Chemistry

Reduction of Cr(VI) to Cr(III) is a potentially important 

mechanism for natural attenuation of Cr in ground water. 

Research on this topic has contributed to identifying potential Cr(VI) reductants and identifying how the rate and extent 

of reaction varies with changes in solution and other conditions. Reduction of Cr(VI) in aqueous solution is promoted 

by low pH and dissolved organic carbon (Stollenwerk and 

Grove, 1985). Dissolved Fe(II) reduces Cr(VI) rapidly over 

a wide range of solution conditions (e.g., Buerge and Hug, 

1997; Eary and Rai, 1988). The rate of reduction by aqueous Fe(II) decreases with decreasing temperature; the effect 

of temperature is more pronounced at near-neutral than at 

low pH (Sedlak and Chan, 1997). Dissolved sulfide (e.g., 

Pettine et al., 1998), sulfur-containing dissolved organic 

compounds (Schroeder and Lee, 1975), and hydrogen 

sulfide gas (Thornton and Amonette, 1999) have been 

shown to reduce Cr(VI). Aqueous reductants like these 

are not likely to be of widespread importance in natural 

attenuation because of the limited extent to which mixing 

occurs during transport; reductants associated with solids 

are likely to be much more important in natural attenuation 

of Cr(VI) (Palmer and Puls, 1994).

Many compounds found in soils and sediments have been 

shown to be capable of reducing Cr(VI). Reduction of Cr(VI) 

by Fe(II) associated with surfaces of oxide minerals (e.g., 

magnetite, ilmenite, and Fe(II)-substituted goethite, White 

and Peterson, 1996; Bidoglio et al., 1993), silicates (Ilton et 

al., 1997; Eary and Rai, 1989), and sulfides (Patterson et al., 

1997) has been reported. Reduction by Fe(II) associated 

with minerals has been observed in the presence of dissolved oxygen in laboratory experiments (Ilton et al., 1997) 

and in field-scale transport experiments (Kent et al., 1994). 

Laboratory experiments have demonstrated that reduction 

of Cr(VI) by soil humic and fulvic substances occurs at an 

appreciable rate at pH 2 but the rate of reduction decreases 

with increasing pH (Wittbrodt and Palmer, 1996), which is 

consistent with reduction of Cr(VI) by synthetic, substituted 

phenols reported by Elovitz and Fish (1994). The rate of 

reduction by soil humic and fulvic substances decreases 

with decreasing temperature in the range 15-55 °C (Wittbrodt and Palmer, 1996). The rate of reduction of Cr(VI) by 

soil organic matter may decrease during the course of the 

reaction as a result of disappearance of more reactive components of the complex soil organic compounds (Wittbrodt 

and Palmer, 1996). Soil organic matter may also promote 

Cr(VI) reduction by enhancing the reductive dissolution 

of Fe(III) to produce Fe(II) (Wittbrodt and Palmer, 1996). 

Reduction of Cr(VI) by low molecular weight aliphatic and 

aromatic organic acids is catalyzed by various titanium-, Al-, 

and Fe(III)-oxides (Deng and Stone, 1996a, b) representative of solids found in soils and sediments.

The reactivity of Fe(II) at mineral surfaces is subject to 

passivation as a result of build up of oxidation products at 

the surface. Reduction of Cr(VI) by magnetite and illmenite 

was inhibited by build-up of Fe(III) oxides on the surface 

resulting from reaction between Fe(II)-containing minerals 

and Cr(VI) or by prolonged exposure of Fe(II) mineral to 

oxygen (Peterson et al., 1997; White, and Peterson, 1996). 

Magnetite collected from anoxic reservoir sediments reduced Cr(VI) rapidly but magnetite collected from an oxic 

soil profile did not reduce Cr(VI), presumably as a result of 

passivation by Fe(III) oxide coatings on the surface (White 

and Peterson, 1996). Passive oxide surface coatings can 

adsorb Cr(VI) but can completely inhibit its reduction (Peterson et al., 1996).

Oxidation of Cr(III) to Cr(VI) is an important process to consider for the long-term performance of a natural attenuation 

alternative. Oxidation of Cr(III) to Cr(VI) has been observed 

in some soil and sediment slurries under oxidizing conditions (Palmer and Puls, 1994; Masscheleyn et al., 1992; 

Bartlett and James, 1979). These studies involved adding 

Cr(III)-containing solutions to soil or sediment slurries. 

Oxidation of naturally occurring Cr(III) to Cr(VI) in soils has 

been reported (Chung et al., 2001), but the form in which 

Cr(III) occurred in the soils and the mechanism(s) by which 

it was oxidized were unknown.

Laboratory experimental studies have provided insight into 

possible mechanisms by which Cr(III) can be oxidized to 

Cr(VI) in soils and sediments. Oxidation of Cr(III) by dissolved O2 in homogeneous solution is too slow to be an 

important process at low temperatures (Nakayama et al., 

1981) but may be appreciable at elevated temperatures 

(Schroeder and Lee, 1975). The potential for solid surfaces 

to catalyze the oxidation of Cr(III) by dissolved oxygen 

has not been adequately studied. Various Mn(IV) and 

Mn(III) oxides can oxidize Cr(III) to Cr(VI) rapidly (Fendorf 

and Zasoski, 1992; Eary and Rai, 1987; Manceau, and 

Charlet, 1992). The reaction is inhibited by precipitation 

of Cr(OH)3 on the Mn oxide surface (Banerjee and Nesbitt, 

1999; Fendorf et al., 1992). At near-neutral pH values, the 

oxidation is not inhibited by Co(II) or Mn(II) but is inhibited 

by Al (Fendorf et al., 1993), which apparently blocks sites 

of oxidation either by surface precipitation or competitive 

adsorption. The oxidation reaction is also inhibited by dissolved organic compounds that form complexes with Cr(III) 

(Johnson and Xyla, 1991; Nakayama et al., 1981). These 

inhibitory effects suggest that oxidation requires contact 

between aqueous Cr(III) and reactive sites at the Mn oxide 

surface. Thus, while Mn oxides are common constituents 

of soils and aquifer sediments and, therefore, potentially 

important oxidants for Cr(III), their ability to oxidize Cr(III) 

will need to be determined on site-specific materials over 

ranges of aquifer chemistry that are relevant to field applications.

Colloidal Transport

There is evidence in natural systems of association of 

chromium with mobile colloidal solids in surface water 

(e.g., Stolpe et al., 2005) and ground water (e.g., Jensen and Christensen, 1999). In general, these colloids 

consist of organic macromolecules, iron oxyhydroxides, 

clay minerals, or sulfide minerals. Increased mobility of 

chromium-bearing colloidal material may result either from 

changes in the surface charge on colloids due to changes 

in subsurface geochemistry (e.g., Grolimund and Borkovec, 
47

2006) or through deflocculation and resuspension of colloidal material through dissolution of cementing agents 

within the aquifer matrix (e.g., Ryan and Gschwend, 1990; 

Ryan and Gschwend, 1992). Both processes would be 

facilitated in aquifers impacted by organic contaminants 

where microbial activity may be stimulated resulting in the 

generation of reducing conditions and/or the production of 

low molecular weight organic compounds that partition to 

fine-grained sediments. Due to its particle reactivity, mobile 

forms of Cr(III) may be attributed, in part, to association 

with mobile colloids. The distance of colloidal transport 

from an impacted zone is uncertain, since colloid stability 

may change significantly during transport from reduced to 

oxidized zones or due to encounter with new sediment surfaces in unimpacted zones. However, identification of this 

transport mechanism within a chromium plume may play 

an important role relative to the management of the source 

zone or the establishment of a monitoring system appropriate for determining the extent of colloidal transport.

Site Characterization

Overview

In most cases the reduction of Cr(VI) to Cr(III) will be the 

process most likely to provide long-term attenuation and 

immobilization of Cr in contaminant plumes and, therefore, 

site characterization should focus on assessing the potential effectiveness of this process. As noted in the previous 

section, Cr oxidation-reduction chemistry is characterized 

by disequilibrium. Therefore, assessments based on equilibrium assumptions, such as Eh or pe determinations, will 

be of little use. Instead, it is recommended that the assessment focus on: 1) identifying solid and solution species 

capable of reducing Cr(VI), 2) quantifying the capacity for 

Cr(VI) reduction, 3) determining the quantity of Cr(VI) that 

could be released over time, 4) assessing possible changes 

in ground water chemistry that could influence the mobilization of soil- and sediment-bound Cr(VI), 5) assessing 

the long-term fate of reductants given possible changes in 

ground-water chemistry, and 6) assessing the potential for 

re-oxidation of Cr(III).

Assessment of Cr Aqueous Speciation

Identification of the mobile form of chromium within the 

aquifer is important relative to the development of a conceptual site model of the process(es) controlling immobilization of this contaminant. Chromium may be mobile as 

a dissolved species or in association with mobile colloidal 

matter. Generally, the anionic Cr(VI) species is the most 

mobile form of dissolved chromium in ground water with 

neutral to basic pH and measurable oxygen (Ball and Izbicki, 

2004), although Cr(III) bound to mobile forms of natural 

organic matter may occur under certain conditions (Li and 

Xue, 2001). The distribution of these potential species in 

ground water provides information relative to the dominant 

process controlling attenuation. Parks et al. (2004) provide 

a review of several published methods for determining chromium speciation in water samples, including those similar to 

Table 5.1 Natural attenuation and mobilization pathways for chromium.

Attenuation Processes Mobilization Processes Characterization Approach

Precipitation of metal 

chromates or precipitation of Cr(III) oxyhydroxide or sulfide

Dissolution of metal chromates due to change in pH; 

dissolution of Cr(III) oxyhydroxide due to acidification; 

dissolution of Cr(III) sulfide 

due to shift from reducing to 

oxidizing conditions.

Evaluation of Cr speciation in ground water. Determination of total Cr in the solid matrix and suspected 

components in chromium-bearing mineral. Evaluation 

of mineral solubility relative to ground-water chemistry 

and published solubility constants. Determination of 

Cr(VI) reductants in ground water and aquifer solids.

Co-precipitation of Cr 

as a trace component in 

oxyhydroxides or sulfides 

of iron or manganese

Dissolution of host oxyhydroxide due to decrease in pH or 

shift from oxidizing to reducing 

conditions; dissolution of host 

sulfide due to shift from reducing to oxidizing conditions.

Evaluation of Cr speciation in ground water. Evaluation 

of host precipitate formation relative to existing groundwater chemistry; determination of host mineral content 

in aquifer sediments via mineralogical characterization. 

Evaluation of Cr solid-phase partitioning using sequential extraction methodologies.

Adsorption of chromate 

to iron oxyhydroxides, 

iron sulfides, or other 

mineral surfaces 

Desorption at high pH for 

oxyhydroxides and sulfides; 

complexation/stabilization in 

the presence of DOC. Reductive dissolution of iron hydroxides or oxidative dissolution of 

iron sulfides.

Evaluation of Cr speciation in the aqueous phase. Determination of total Cr in the solid matrix. Evaluation of 

Cr solid-phase partitioning using sequential extraction 

methodologies. Batch and column testing to determine 

Cr uptake capacity of site-specific aquifer materials 

with variable geochemical conditions.
48

published EPA methods (Table 5.2). These authors stress 

the importance of sample preservation and processing on 

the determination of total Cr and Cr(VI) in water samples, 

particularly in samples that may contain elevated concentrations of colloidal iron oxyhydroxides that may pass 

through a 0.45 μm filter or precipitate in-situ during storage 

of improperly preserved samples.

Differentiation of Cr(III) and Cr(VI) species is typically 

determined via two approaches: 1) chemical separation 

of the two species by ion-exchange or coprecipitation followed by direct determination of total dissolved chromium 

and 2) selective determination of Cr(VI) by colorimetry or 

differential pulse polarography. Colorimetric methods for 

direct determination of Cr(VI) exist for field measurements 

and may be employed if there are concerns about the ability to preserve chromium speciation prior to analysis in a 

laboratory setting. A recent study provides a useful comparison of the relative performance of a field colorimetric 

method, a field method for species separation followed by 

laboratory analysis, and EPA Method 218.6 for laboratory 

separation and quantification of Cr(III) and Cr(VI) by ion 

chromatography (e.g., Ball and McCleskey, 2003a and 

2003b). Ultimately, the method employed for speciation 

of aqueous chromium will depend on required analytical 

detection limits and sample preservation requirements, so 

it is recommended that analytical performance be demonstrated on a site-specific basis.

It has also been observed that chromium may be leached 

from certain grades of stainless steel well casing/screen 

materials under chemical conditions that may be encountered in contaminant plumes. Two published studies have 

provided detailed evaluation of the extent of chromium 

(and nickel) leaching that may occur for type 304 or 316 

stainless steel screens (Hewitt, 1994; Oakley and Korte, 

1996). Oakley and Korte (1996) provide a site-specific 

example of how elevated chromium concentrations derived 

from continuous leaching of well screen materials may be 

falsely identified as a component of a ground-water plume. 

This suggests that careful consideration should be given to 

the types of well screen materials and sampling protocols 

employed relative to the assessment of potential contaminants of concern within a plume.

Identifying and Quantifying Cr(VI) 

Reductants

Ground-water characteristics of direct interest to assessment of Cr(VI) reduction potential include pH and concentrations of dissolved oxygen, Fe(II), sulfide, and organic 

carbon. Ground-water pH can influence Cr(VI) reduction in 

two ways. First, from a thermodynamic perspective, the affinity (Morgan, 1967) of many oxidation-reduction reactions 

involving the Cr(VI)-Cr(III) couple depends on pH. Second, 

many of the rates of oxidation-reduction reactions depend 

on pH. The presence of high concentrations of dissolved 

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

Method Name and Number Source

Chromium, Hexavalent - Chelation Extraction; 

7197

a SW-846 Ch 3.3; 

http://www.epa.gov/epaoswer/hazwaste/test/pdfs/7197.pdf

Chromium, Hexavalent - Colorimetric Method; 

7196A

a SW-846 Ch 3.3; 

http://www.epa.gov/epaoswer/hazwaste/test/pdfs/7196a.pdf

Chromium, Hexavalent - Coprecipitation Method; 

7195

a SW-846 Ch 3.3; 

http://www.epa.gov/epaoswer/hazwaste/test/pdfs/7195.pdf

Chromium, Hexavalent - Differential Pulse Polarography; 7198

a SW-846 Ch 3.3 

http://www.epa.gov/sw-846/pdfs/7198.pdf

Determination of Hexavalent Chromium in Drinking Water, Groundwater, and Industrial Wastewater 

Effluents by Ion Chromatography, 7199

a SW-846 Ch 3.3; 

http://www.epa.gov/sw-846/pdfs/7199.pdf

Metals and Trace Elements by ICP/Atomic Emission Spectrometry; 200.7 Rev. 4.4 b EPA/600/R-94/111 (NTIS Order Number PB95-125472)

Trace Elements by ICP/Mass Spectrometry; 200.8 

Rev. 5.4 b EPA/600/R-94/111 (NTIS Order Number PB95-125472)

Trace Elements by Stabilized Temperature Graphite Furnace AA Spectrometry; 200.9 Rev. 2.2 b EPA/600/R-94/111 (NTIS Order Number PB95-125472)

a EPA SW-846 Ch 3.3, Test Methods for Evaluating Solid Waste, Physical/Chemical Methods 

(http://www.epa.gov/epaoswer/hazwaste/test/main.htm)

b Methods for the Determination of Metals in Environmental Samples-Supplement I, EPA/600/R-94/111, May 1994. Available at 

National Technical Information Service, PB95-125472. (http://www.ntis.gov/products/types/publications.asp?loc=4-4-4)
49

oxygen is indicative of conditions unfavorable for Cr(VI) 

reduction. In contrast, low but detectable dissolved oxygen concentrations can indicate mildly reducing conditions 

favorable for the accumulation of Cr(VI)-reductants in the 

sediments. Detectable concentrations of Fe(II) or sulfide 

are indicative of conditions under which Cr(VI) reduction 

should occur rapidly. High concentrations of dissolved 

organic carbon could indicate that organic compounds 

capable of reducing Cr(VI), either directly or catalyzed by 

mineral surfaces, may be present.

Detection of Cr(VI) reductants in ground water signals 

conditions favorable for natural attenuation, but most of 

the Cr(VI) reduction capacity will likely reside in the sediments (Palmer and Puls, 1994). Potentially important Cr(VI) 

reductants include sulfides, Fe(II) minerals, and organic 

matter. Even in the absence of detectable concentrations of 

reductants in ground water, Cr(VI) reductants may be present in the sediments. For example, sediment-bound Fe(II) 

can provide a significant capacity for natural attenuation 

of Cr(VI) even in the absence of detectable concentrations 

of Fe(II) or sulfide and the presence of low but detectable 

concentrations of dissolved oxygen (Kent et al., 1994).

In most cases the capacity for Cr(VI) reduction by soils or 

aquifer sediments will have to be determined experimentally using site-specific materials. Laboratory experiments 

can be used to assess the Cr(VI) reduction capacity and 

provide insights into reduction rates by constituents of the 

sediments. However, it should be noted that exposure to atmospheric oxygen can greatly decrease the Cr(VI) reductive 

capacity of sediments (e.g., Anderson et al., 1994). Also, 

rates of reduction measured in laboratory experiments may 

exceed those achieved during transport. For example, the 

rate of reduction of Cr(VI) determined in laboratory batch 

experiments greatly exceeded that during transport in a 

sand and gravel aquifer because, in the field, the rate of 

reduction was limited by the rate of mass transfer across 

sediment layers (Friedly et al., 1995).

Use of Cr Isotopes to Determine Reduction 

Rates

Recent work suggests that in-situ Cr(VI) reduction rates 

may be determined through the assessment of chromium 

stable isotope ratios (53Cr/52Cr) in ground water (Ellis et 

al., 2002). Cr(VI) reduction causes an enrichment of the 

lighter isotopes in the reduced product at any given instant. 

Faster reduction of lighter isotopes of Cr(VI) as compared 

to heavier isotopes resulted in measurable changes in 

the isotopic composition of dissolved Cr(VI) along flow 

paths where reduction was occurring. Thus, the extent 

of Cr(VI) reduction could be assessed along a flow path 

provided the isotopic composition of the initial source is 

known. Additional work indicates that sorption reactions 

occurring within the fringe of contaminant plumes may 

contribute to the observed isotope fractionation, but this 

work also indicated that sorption would exert little impact 

in more mature portions of the plume where sorption may 

play a less dominant role (Ellis et al., 2004). Theoretical 

calculations support the mechanism of proposed isotopic 

fractionation between Cr(VI) and Cr(III) (Schauble et al., 

2004; Ottonello and Zuccolini, 2005a; Ottonello and Zuccolini, 2005b) suggesting that characterization of chromium 

stable isotope ratios may prove a useful tool for assessing 

in-situ rates of Cr(VI) reduction.

Assessment of Cr Solid-Phase Speciation

Various chemical extraction procedures have been proposed for quantifying soil- and sediment-bound Cr(VI) 

contamination. Studies in which the effectiveness of 

different extraction techniques were compared showed 

that a method involving a heated carbonate-hydroxide 

solution was the most effective for recovering soluble and 

insoluble forms of Cr(VI) (James et al., 1995; Vitale et al., 

1997). For sparingly soluble Cr(VI) salts, the method was 

effective at quantifying PbCrO4 but only partially effective 

at quantifying BaCrO4, which may be an important form 

of Cr(VI) at some sites (Palmer and Puls, 1994). This 

method has been adopted by the USEPA (Method 3060A; 

http://www.epa.gov/epaoswer/hazwaste/test/pdfs/3060a.

pdf) for determination of Cr(VI) in contaminated soils and 

waste materials. It should be noted that Vitale et al. (1997) 

demonstrated that this method may not be appropriate for 

solid materials containing sulfide precipitates that reduce 

Cr(VI) to Cr(III) upon dissolution during extraction. Figure 

5.1 suggests that a mildly acidic (pH 3-4) acid leach in the 

presence of sufficient sulfate to precipitate Ba as BaSO4

might effectively dissolve BaCrO4, but this has not been 

tested. Very low pH values should be avoided because of 

the potential for causing Fe(III) concentrations to increase 

sufficiently to precipitate chromian jarosite (Figure 5.2). At 

least one potential artifact that could result in overestimating 

sediment-bound Cr(VI) has been noted. Storing soils so 

that they are partially saturated (with water) may result in 

oxidation of native Cr(III) to Cr(VI); drying the soils prior to 

storage apparently eliminates this possible artifact (Chung 

et al., 2001). Korolczuk (2000) has examined an adaptation 

of the heated carbonate-hydroxide extraction method that 

employs voltammetric determination of extracted Cr(VI) 

in organic-rich soils containing predominantly Cr(III) solid 

species. This approach was more reliable for Cr(VI) detection than use of the colorimetric procedure described 

in Method 3060A.

Various extraction schemes have been employed to determine the chemical speciation of the total chromium 

content of contaminated soils or sediments. The intent of 

this solid-phase characterization approach is generally to 

identify the solid-phase partitioning reactions that govern 

chromium mobility in the subsurface. However, as noted 

by Szulczewski et al., (1997), there are limitations to the 

selectivity of various solutions proposed to extract chromium 

from labile and recalcitrant pools of solid species within 

soils and sediments. These authors as well as Wilkin et al. 

(2005) have made use of x-ray absorption spectroscopy to 

assist in defining the chemical speciation of chromium in 

contaminated soils and aquifer sediments, but little work 

has been conducted to date to better design and assess 

the accuracy of chemical extractions to determine the 

chemical speciation of chromium in solid matrices. Thus, 
50

currently Method 3060A provides the most thoroughly 

tested approach for assessing the presence of Cr(VI) in 

soils and sediment.

 Long-Term Stability and Capacity

Changing aquifer chemical conditions will influence the 

processes that control the mobility and natural attenuation of Cr(VI). Changes in ground-water pH and dissolved 

salt concentrations can change the solubility of Cr(VI) 

compounds (Figures 5.1 and 5.2) and the extent to which 

Cr(VI) adsorbs on the sediments (Figure 5.3). Acidification 

of natural attenuation zones can cause the Cr(VI) reductive 

capacity of the sediments to decrease (e.g., Anderson et 

al., 1994) or to increase (e.g., Wittbrodt and Palmer, 1996; 

Stollenwerk Grove, 1985). Acidification can also increase 

the mobility of Cr(III) (e.g., Walter et al., 1994). Invasion of 

oxygen into reducing zones can result in oxidation of Cr(VI)-

reductants. Establishment of completely oxic conditions 

may result in production of Mn oxides or other compounds 

capable of re-oxidizing Cr(III) previously sequestered in 

the sediments.

Laboratory experimental studies conducted during the site 

assessment phase will provide insight into the impact of 

anticipated changes in aquifer chemistry on many of these 

processes. However, the long-term performance of natural 

attenuation under changing aquifer chemical conditions will 

be difficult to predict with certainty. Therefore, long-term 

monitoring of chemical conditions in and around zones in 

which attenuation processes are active is recommended. It 

is recommended that long-term monitoring include the critical water quality parameters identified above. If chemical 

conditions become increasingly oxidizing, then it is recommended that long-term monitoring be supplemented with 

periodic assessment of the Cr(VI) reduction capacity of 

the sediments.

Tiered Analysis

Determination of the viability of chromium remediation in 

ground water via monitored natural attenuation will depend 

upon proper assessment of contaminant loading to the 

aquifer and prevailing geochemistry and mineralogy within 

the contaminant plume and the down gradient zone prior 

to the point(s) of compliance. The goal of site assessment 

is to demonstrate the process(es) controlling chromium 

sequestration onto aquifer solids and the long-term stability of solid phase chromium as a function of existing and 

anticipated ground-water chemistry. The following tiered 

analysis structure for site characterization provides an approach to evaluate candidate sites and define the potential 

limitations of MNA as part of a remedy for ground-water 

cleanup.

Tier I. Site characterization under Tier I will involve demonstration that the plume is static or shrinking, has not 

reached compliance boundaries, and does not impact 

existing water supplies. Once this is established through 

ground-water characterization, evidence is collected to 

demonstrate Cr partitioning to aquifer solids within the 

plume. Rapid movement of contaminants along preferred 

flow paths in the unsaturated and saturated zones can by 

induced by hydrologic events such as heavy rains (e.g., 

McCarthy et al., 1998; Camobreco et al., 1996). It will be 

important to determine that such hydrogeologic features do 

not result in contaminants bypassing zones where natural 

attenuation is occurring. If natural attenuation processes 

are active throughout the plume, then there should be an 

observed increase in solid phase concentrations within 

regions of the plume with higher aqueous concentrations, 

e.g., near the source term. This field partitioning data 

may be supplemented by geochemical modeling that 

incorporates measured water chemistry (e.g., pH, Eh, 

and major ion chemistry) throughout the plume to assess 

the potential for solubility control by metal chromates or 

chromium oxyhydroxides or sulfides. Since identification 

of the chemical speciation of chromium in water samples 

and aquifer sediments is critical towards determining the 

attenuation mechanism(s), it is recommended that precautions be taken to preserve chromium speciation during collection, preservation, and processing of collected samples 

(See recommendations and application of these methods 

in Wilkin et al., 2002 and Ford et al., 2005). Identification 

of active sequestration to prevent chromium migration in 

ground-water provides justification for proceeding to Tier 

II characterization efforts.

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

of attenuation are determined. Estimates of a site attenuation rate(s) can be assessed via a well transect along the 

ground-water flow path. This analysis may be supplemented 

by analysis of the distribution of Cr isotopes in ground-water 

along the flow path. In addition, time-series data may be 

collected at one or more monitoring points within the plume. 

This information will allow assessment of the relative timescales for contaminant immobilization and fluid transport 

and determination of whether remediation objectives can 

be met within the required regulatory time frame. In addition, the mechanism(s) for attenuation need to be identified 

under this stage of site characterization. This effort may 

require determination of the chemical speciation of aqueous and solid phase chromium, which may be approached 

according to the following scheme:

1) Determination of chromium solution speciation 

via direct analytical measurements in combination with speciation calculations based on 

characterized ground-water chemistry (e.g., 

Jensen and Christensen, 1999);

2) Calculation of the saturation state of ground 

water relative to measured aqueous chemistry 

(e.g., Wilkin et al., 2005) complimented by the 

possible isolation of discrete Cr mineral phases 

via density separations (or other schemes) in 

regions of the aquifer with highest solid phase 

concentrations;

3) Determination of aquifer mineralogy to determine the relative abundance of components 

with documented capacity for Cr sorption (e.g., 

Amonette, 2002);
51

4) Identification of chromium association(s) with 

the various solid phase components of aquifer 

solids through combination of chemical extractions with microscopic/spectroscopic confirmation of phase associations, and;

5) Demonstration of concurrence between the site 

conceptual model and mathematical model(s) 

that describe arsenic removal mechanism(s).

It is recommended that identification of chromium chemical 

speciation in aqueous and solid matrices be conducted using samples collected in a manner that preserves the in-situ 

mineralogy and speciation of chromium. The demonstration of concurrence between conceptual and mathematical 

models describing arsenic transport will entail development 

of site-specific parameterization of the chemical processes 

controlling chromium solid-phase partitioning.

Tier III. Once the partitioning mechanism(s) have been 

identified for the site, the subsequent characterization effort under Tier III will involve determination of the stability 

of immobilized Cr and the capacity of the aquifer to sustain continued uptake. The impact of potential hydrologic 

changes, such as a shift in flow direction caused by the 

onset of pumping at nearby sites, needs to be determined. 

A well-constrained ground-water flow model of the site 

should be capable of evaluating the impact of such changes. 

It is recommended that the stability of immobilized Cr be 

tested based on the anticipated evolution of ground-water 

chemistry concurrent with plume shrinkage. For example, 

changes in ground-water redox potential due to oxygen intrusion or the occurrence of anaerobic microbial processes 

can exert a significant influence on Cr partitioning to iron 

sulfides or iron oxyhydroxides, respectively. Therefore, it is 

recommended that sediment leach tests be conducted to 

characterize the magnitude of Cr mobilization as a function 

of ground-water redox chemistry representative of existing 

and anticipated site conditions. This may involve the use of 

microcosm tests that stimulate in-situ microbial populations 

toward the development of redox conditions considered 

deleterious for continued Cr immobilization. It is recommended that the capacity for Cr uptake onto aquifer solids 

be determined relative to the specific mechanism(s) identified in Tier II. For example, if site characterization under 

Tier II indicated that co-precipitation of Cr with iron sulfide 

occurs due to microbial degradation of organic compounds 

coupled with sulfate reduction within the aquifer, it is recommended that the mass distribution of organic carbon, 

sulfate and ferrous iron to support this reaction within the 

aquifer be determined. This site-specific capacity can then 

be compared to Cr mass loading within the plume in order 

to assess the longevity of the natural attenuation process. 

If site-specific tests demonstrate the stability of immobilized 

Cr and sufficient capacity within the aquifer to sustain Cr 

attenuation, then the site characterization effort can progress to Tier IV. For cases where contaminant stability is 

sufficient but aquifer capacity is insufficient for capture of 

the entire plume, then a determination of the benefits of 

contaminant source reduction may be necessary.

Tier IV. Finally, under Tier IV a monitoring plan is established along with contingency plans in the event of MNA 

failure. It is recommended that the monitoring plan be 

designed to establish both continued plume stability and to 

identify changes in ground-water chemistry that may lead 

to re-mobilization of attenuated Cr. The specific chemical 

parameters to be monitored will include those identified 

under Tier III that may halt Cr partitioning to aquifer sediments and/or result in solubilization of either discrete Cr precipitates or aquifer minerals that sequester Cr from ground 

water. For example, solution phase parameters that could 

alter Cr(III) precipitation include decreases in ground-water 

pH. Similarly, increases in the concentration of competing 

anions, such as phosphate, could lead to re-mobilization 

of adsorbed Cr(VI). In contrast, the concentration of dissolved iron or sulfate may indicate the dissolution of an 

important sorptive phase within the aquifer (e.g., reductive 

dissolution of iron oxyhydroxides or oxidative dissolution of 

sulfides). Changes in these parameters may occur prior to 

observed changes in solution Cr and, thus, serve as monitoring triggers for potential MNA failure. In this instance, 

a contingency plan can be implemented that incorporates 

strategies to arrest possible plume expansion beyond compliance boundaries. Possible strategies to prevent plume 

expansion include pump and treat operations, installation 

of reactive barriers to enhance uptake capacity perpendicular to the direction of plume advance, or enhancement 

of natural attenuation processes within the aquifer through 

the injection of soluble reactive components.

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57

Arsenic

Robert G. Ford, Douglas B. Kent, Richard T. Wilkin

Occurrence and Distribution

The types of arsenic wastes at Superfund sites include 

byproducts of wood preserving, arsenic waste disposal, 

pesticide production and application, and mining (Table 1 in 

USEPA, 1997). Mandal and Suzuki (2002) provide a review 

of anthropogenic sources of arsenic in terrestrial systems. 

The sources of arsenic in soils are diverse, and they may 

be derived from organic or inorganic forms. For example, 

Foster et al. (1998) examined the chemical speciation of 

arsenic in mine wastes to show that the waste material 

was chemically complex and composed of a combination 

of arsenic in native mineral forms and sorbed to common 

weathering products. Jackson and Bertsch (2001) examined the types of aqueous chemical species of arsenic 

released from leaching of organo-arsenic materials used in 

agricultural applications. The aqueous species of arsenic 

released during the weathering of arsenic compounds in 

anthropogenic sources may be distinct from the subsequent 

speciation of arsenic in ground water. Leached species 

may be unstable in the ground-water environment, but 

serve as intermediates that control the overall release of 

arsenic from source zones. 

Arsenic in soil and ground water is derived from natural 

and/or anthropogenic sources (USEPA, 1997). Natural sources of arsenic are derived from a wide array of 

geologic materials, including igneous, metamorphic and 

sedimentary rocks (Korte and Fernando, 1991; Welch et 

al., 2000; Smedley and Kinniburgh, 2002). Arsenic may 

subsequently be accumulated during secondary mineral 

formation in overburden materials and soils. In contrast, 

anthropogenic sources are typically derived from the land 

application of arsenical pesticides and herbicides and 

from disposal of arsenic-bearing wastes generated during 

processing of ore materials for production of commercial 

products (USEPA, 1995).

Natural sources of arsenic may also pose a health risk in 

ground water due to changes in site geochemistry that 

promote mobilization. An example of this scenario is the 

mobilization of naturally occurring arsenic during transport 

of landfill leachate through the ground-water aquifer (Hounslow, 1980). In this case, arsenic is mobilized via desorption 

induced by competitive interactions with dissolved constituents or due to dissolution of host mineral phases as a result 

of microbial degradation of organic compounds within the 

contaminant plume. In order to evaluate the potential for 

down gradient attenuation of mobilized arsenic, it is recommended that site characterization be sufficient to identify 

the potential sinks for arsenic and the chemical conditions 

that would result in uptake onto aquifer solids.

Plume Characteristics

Hounslow (1980) and Smedley and Kinniburgh (2002) provide an assessment of geochemical triggers that may lead 

to arsenic mobilization in subsurface systems. In general 

these include: 1) desorption at high pH under oxidizing 

conditions and/or due to the influx of dissolved ions that 

compete for sorption sites on aquifer minerals, 2) desorption/dissolution due to a change to a reducing chemical 

environment, and 3) mineral dissolution. The first process is 

a result of the influx of dissolved constituents that compete 

for or displace arsenic adsorbed to mineral surfaces without 

a concomitant change in sorbent stability. The second process is a result of a change in the ground-water chemistry 

to a condition under which the sorbent material is no longer 

stable. This may be a result of a change in redox, pH or 

other factors leading to dissolution of the sorbent phase. 

The third process is a result of a shift in arsenic mineral 

stability due to a change in ground-water chemistry. In this 

instance, arsenic may represent a major or minor component within the solid, e.g., orpiment or arsenic-rich pyrite, 

respectively. While site-specific mechanisms may differ 

between naturally occurring versus contaminant sources 

of arsenic, these general arsenic mobilization processes 

apply under both scenarios.

Remedial Technologies

Technologies for the remediation of arsenic in soil involve 

treatments to effect containment, immobilization, or separation/concentration within the solid matrix. The separation/concentration process is followed by some secondary 

immobilization treatment (USEPA, 1995; USEPA, 1997). 

Technologies for the treatment of arsenic in ground water 

are based on ex-situ or in-situ approaches. Pump-andtreat technologies make use of processes common to 

water and wastewater treatment for removal of dissolved 

arsenic (USEPA, 2002). In-situ treatment technologies are 

less common, but there is emerging research based on 

the application of permeable reactive barriers for arsenic 

removal from ground water. This technology is based on 

installation of reactive solid material into the subsurface to 

intercept and treat the contaminant plume (Lackovic et al., 

2000; Su and Puls, 2001; Cheng et al., 2005).
58

Regulatory Aspects

Arsenic is the second most common contaminant of 

concern (COC) for sites listed on the Superfund National 

Priorities List (USEPA, 2002). Of the 1209 sites on the 

National Priorities List for which a Record of Decision has 

been signed, there are 380 and 372 sites with arsenic as a 

COC in ground water and soil, respectively. The maximum 

contaminant level for arsenic in drinking water was recently 

revised from 0.05 mg L-1 to 0.01 mg L-1 (USEPA, 2001; 

USEPA, 2006a; http://www.epa.gov/waterscience/criteria/

drinking/dwstandards.pdf). This may impact site-specific 

cleanup goals at hazardous waste sites for locations where 

ground water is the primary source of potable water. For 

non-potable water sources, ambient water quality criteria 

(AWQC) that are protective of aquatic life may serve as 

alternative cleanup goals. For arsenic, current statutes list 

both acute and chronic criteria for arsenic in fresh waters as 

0.34 mg L-1 and 0.15 mg L-1, respectively (USEPA, 2006b; 

http://www.epa.gov/waterscience/criteria/nrwqc-2006.pdf). 

Contaminant cleanup goals vary widely for arsenic in soil 

depending on the use of total or leachable metals as the 

cleanup criterion and the potential impact and intended use 

of ground water supplies (USEPA, 2001). 

Geochemistry and Attenuation Processes

Aqueous Speciation

In oxidizing environments, the predominant form of arsenic 

in solution is arsenate, HnAsO4

n-3. The arsenate oxyanion 

may be protonated to various degrees as a function of 

pH, but it is commonly present as the negatively-charged 

H2AsO4

1- or HAsO42- within the pH range of natural waters 

(Cullen and Reimer, 1989; Nordstrom and Archer, 2003). 

However, the neutral (H3AsO4

0

) or the fully deprotonated 

form (AsO4

3-) may exist at acid or alkaline pH extremes that 

may occur at contaminated sites.

The speciation of arsenic may be more complex in reducing environments. In general, arsenite, Hn

AsO3

n-3, 

is the predominant arsenic species, but mononuclear 

(HxAsOySz

(3+x–2y–2z)) and polynuclear thioarsenic species 

may also form in sulfate-reducing zones in which iron is 

depleted (Clarke and Helz, 2000; Wilkin et al., 2003; Bostick 

et al., 2005; Hollibaugh et al., 2005). The distribution of 

various reduced arsenic species under relevant geochemical conditions for an arsenic contaminated site is shown in 

Figure 6.1. These model results suggest that arsenite would 

predominate under most reducing environments. However, 

thioarsenic species may become significant in sulfate-reducing zones that are depleted in iron. These conditions may 

be encountered in organic-rich ground water derived from 

contaminated sites (Vroblesky and Chapelle, 1994). The 

aqueous concentration of thioarsenic species may remain 

high in sulfide-dominated systems due to the solubility of 

orpiment at circumneutral pH (Webster, 1990). The detection of the species controlling aqueous arsenic speciation 

in reducing environments may be complicated by the need 

to employ different sample preservation techniques in the 

absence or presence of aqueous sulfide.

Figure 6.1 The distribution of arsenite and thioarsenic 

species in a reducing ground water. The 

sum of all possible thioarsenic species is 

shown for simplicity. The dashed line indicates the current MCL for arsenic in drinking 

water (10 ppb). Model results were generated using MinteqA2 following modification 

of the database to include mononuclear 

thioarsenic equilibrium expressions (Wilkin 

et al., 2003). Calculations were carried 

out with a background electrolyte of 0.1 M 

NaNO3, and mackinawite (FeS) was allowed 

to precipitate in systems with Fe2+.

Solubility

The precipitation of pure phase arsenic minerals is not likely 

in ground-water systems. Examination of solubility trends 

for phases such as As2O5, As2O3, and As2S3 indicates that 

these phases would form only under extreme conditions. 

Metal arsenates are relatively insoluble (e.g., Ca3(AsO4)

2

and Ba3(AsO4)

2

), but formation of these phases is limited 

by the typically lower concentrations of dissolved metals 

and arsenic encountered in ground water downgradient from 

contaminant source areas. The stability fields of scorodite 

[Fe(III)AsO4·2(H2O)] and possible arsenic sulfides are shown 

in Figure 6.2 as a function of pH and Eh for concentrations 

of total dissolved As, Fe, and S that may be encountered 

within a contaminant plume. There is also recent evidence 

for the microbially-mediated formation of ferric/ferrous iron 

arsenites under certain geochemical environments (Morin 

et al., 2003). Under reducing conditions, the stability fields 

of pure arsenic sulfides are quite narrow for relatively high 

total concentrations of arsenic in ground water, although 

site-specific conditions may support precipitation of sulfides 

(e.g., O’Day et al., 2004). In addition, while the precipitation 

of scorodite seems feasible, this process is typically limited 

by the consumption of dissolved iron during the formation 

of iron oxides or iron sulfides. Thus, the immobilization of 

arsenic via precipitation of pure As-bearing mineral phases 

is typically limited. However, coprecipitation with other 
59

common soil/sediment minerals such as iron oxides and 

iron sulfides is a likely sink for arsenic in ground water. For 

this process, arsenic uptake is concurrent with formation of 

metastable phases of Fe-bearing minerals that can occur 

proximate to transitions from oxidizing-to-reducing conditions (or vice versa) within the subsurface (e.g., Ford et 

al., 2005; Ford et al., 2006; Wilkin and Ford, 2006). This 

process is generally distinct from sorption to pre-existing mineral phases in that arsenic may subsequently be 

transferred to more stable mineral phases over time (e.g., 

Moore et al., 1988; Ford, 2002; Lowers et al., 2007). Given 

the uncertainty in the rates of transformation of metastable 

minerals to more stable forms, it is currently not possible to 

reliably estimate the importance of this stabilization process 

in ground-water systems. However, it is recommended that 

this process be considered in the context of the dynamics 

of site-specific geochemistry and the required timeframe 

for attainment of remedial goals. 

Adsorption

The primary forms of inorganic arsenic in both oxidizing 

and reducing ground water are oxyanions or thiooxyanions 

(Ferguson and Gavis, 1972; Wilkin et al., 2003; Bostick et 

al., 2005). Adsorption of arsenic species at mineral surfaces 

occurs as a result of a set of chemical reactions between 

aqueous species and surface sites (Dzombak and Morel 

1990; Davis and Kent, 1990). Adsorption of both As(III) and 

As(V) onto most minerals exhibits a strong pH dependence 

because: 1) most adsorption reactions between As(III) and 

As(V) and mineral surface sites have H+ as a reactant, 2) 

arsenic speciation varies with pH (Figure 6.2), and 3) the 

electrostatic contribution to the free energy of adsorption 

of As species onto most minerals varies with pH. In addition to pH, the extent to which As(III) or As(V) adsorbs at 

mineral surfaces will be influenced by the concentrations 

of other anions, which compete for surface sites, and cations, the adsorption of which can influence the electrostatic 

contribution to anion adsorption.

Arsenic sorption has been demonstrated for a wide range 

of minerals common to soils and sediments with iron oxides 

and sulfides appearing to play a dominant role in oxidizing 

and reducing environments, respectively (Goldberg and 

Glaubig, 1988; de Vitre et al., 1991; Morse, 1994; McNeill 

and Edwards, 1997; Manning et al., 1998; Chiu and Hering, 2000; Wolthers et al., 2005). The relative distribution 

of inorganic species of arsenic in relation to the distribution 

of potential Fe-bearing minerals as a function of pH and 

Eh in the absence of significant sulfate reduction is shown 

in Figure 6.3. For near-neutral pH under oxidizing conditions, sorption of arsenate onto iron oxides such as goethite 

(α-FeOOH) would likely predominate. Under iron-reducing 

conditions, sorption of arsenate or arsenite to reduced 

Fe-bearing minerals such as siderite (FeCO3

), magnetite 

[Fe(III)2Fe(II)O4], or green rust phases [Fe(II)6Fe(III)2(OH)18

· 4(H2O)] may predominate. However, arsenic may remain 

mobile in Fe-reducing systems in which reduced Fe-bearing 

minerals do not precipitate, resulting in a loss of sorption 

capacity within the aquifer sediment (Swartz et al., 2005; 

Polizzotto et al., 2005; Polizzotto et al., 2006). For reduced 

systems in which iron- and sulfate-reduction processes 

are significant, arsenic may sorb to iron sulfides such as 

pyrite (Figure 6.4). Sorption of arsenic to iron (or other 

metal sulfides) may be the dominant mechanism for arsenic mobilization under sulfate-reducing conditions that 

are undersaturated with respect to precipitation of a pure 

arsenic sulfide (Wilkin and Ford, 2006).

The extent to which inorganic arsenic will partition to mineral 

surfaces will be governed by competition with other anions 

in solution and the net surface charge that develops as a 

function of site-specific geochemistry. There are several 

commonly occurring anions in natural waters that could 

compete with arsenic sorption to mineral surfaces. These 

competitive reactions will be active for all arsenic aqueous 

species in oxidized and reduced systems. There is evidence to support that arsenic desorption may occur due 

to competition with other inorganic and organic anions in 

solution or due to increases in ground-water pH (Peryea, 

1991; Jackson and Miller, 2000; Meng et al., 2000; Redman et al., 2002).

Redox Chemistry

Microbial interactions in aqueous systems can impact arsenic mobility via direct and indirect mechanisms. Experimental evidence suggests that the speciation of arsenic can 

be directly influenced via microbially-mediated reductionoxidation (redox) reactions (Ahmann et al., 1994; Zobrist 

et al., 2000). Common outcomes of these interactions are 

1) the conversion of inorganic arsenic between its oxidized 

or reduced oxyanionic forms and 2) methylation-demethylation of arsenic (Cullen and Reimer, 1989; Anderson and 

Bruland, 1991; Maeda, 1994). The extent to which these 

Figure 6.2 Eh-pH diagram for arsenic at 25°C. System 

As-H2O-Fe-S, with ΣAs=10-5 (750 μg L-1), 

ΣFe=10-4 (5.6 mg L-1), and ΣS=10-3

(32 mg L-1

). Solid stability fields are shaded 

blue. Orpiment(a) and Realgar(a) are poorly 

crystalline forms of As2S3 and AsS, respectively. Arsenopyrite and crystalline forms of 

As2S3 and AsS are suppressed.
60

Figure 6.3 Eh-pH diagrams for arsenic and iron at 25°C for iron-reducing systems. These paired diagrams show the 

relative distribution of potentially adsorbing arsenic species (left) relative to representative types of Febearing sorbents (right) that are predicted to occur as a function of Eh and pH. (a) System As-H2

O, with 

ΣAs contoured from 10-8 to 10-4; region for elemental arsenic is shaded gray. (b) System Fe-C-H2O (no 

sulfur) with ΣFe contoured from 10-2 to 10-6 and ΣC=10-3; Hematite and Wüstite are suppressed.

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

paired diagrams show the relative distribution of potentially adsorbing arsenic species (left) relative to 

representative types of Fe-bearing sorbents (right) that are predicted to occur as a function of Eh and pH. 

(a) System As-S-H2O, with ΣAs=10-5 and ΣS=10-3; all solids suppressed to show stability fields for the 

aqueous species. (b) System Fe-C-S-H2O with ΣFe=10-4, ΣC=10-3, and ΣS=10-3; Hematite is suppressed.
61

processes influence on arsenic mobility in intermediate and 

deep ground-water systems is not well known. However, 

these processes could play a more significant role in shallow 

systems with zones of ground-water discharge-recharge 

where biological activity may be more pronounced. More 

commonly, microbial interactions govern arsenic mobility 

by controlling the redox chemistry of major elements with 

which arsenic is associated, e.g., iron, sulfur, and carbon 

(Moore, 1994; Harrington et al., 1998; McGeehan et al., 

1998; Jones et al., 2000). The development of iron- and/or 

sulfate-reducing zones in ground-water contaminant plumes 

governs the precipitation-dissolution of iron- and sulfurbearing minerals. These reactions influence arsenic mobility 

through production or loss of sorptive material within the 

aquifer (Ford, 2005).

Recent studies have increased awareness of the role 

of microorganisms in catalyzing the oxidation of As(III). 

Chemoautotrophic microorganisms that can oxidize As(III) 

to As(V) using oxygen have been isolated from arsenopyrite-bearing rock (Santini et al., 2000). Chemoautotrophic 

microorganisms that can oxidize As(III) under anaerobic 

conditions using nitrate, nitrite, and, possibly, Fe(III) have 

also been identified (Oremland et al., 2002). Laboratory 

experiments have shown that these microorganisms can 

oxidize millimolar concentrations of As(III) over a period of 

several hours under both aerobic and anaerobic conditions. 

Arsenic-oxidizing chemoautotrophic anaerobes were shown 

to be responsible for maintaining As(V) concentrations up 

to 160 uM in the anoxic hypolimnion of a lake whose sediments were heavily contaminated with As from industrial 

activities (Senn and Hemond, 2002). The dominance of 

As(V) over As(III) in other anoxic water bodies underscores 

the potential importance of microbial As(III) oxidation (Senn 

and Hemond, 2002).

Microbial reduction of As(V) to As(III) is likely to be important 

in some systems. Anaerobes whose dominant metabolic 

process involves coupling As(V) reduction to oxidation 

of organic compounds or hydrogen have been isolated 

from natural and polluted water or sediments (Ahmann 

et al., 1994; Hoeft et al., 2002). Laboratory studies have 

shown that these organisms can reduce As(V) adsorbed 

to hydrous ferric or aluminum oxides (Zobrist et al., 2000) 

or precipitated with Fe(III) or Fe(II) (Ahmann et al., 1997). 

Field studies have demonstrated the importance of these 

microorganisms in reducing As(V) to As(III) associated 

with As-contaminated lake sediments (Ahmann et al., 

1997); As(V) metabolism under anaerobic conditions was 

shown to be responsible for at least 14 percent of the carbon mineralization in hypersaline Mono Lake, California, 

where arsenic concentrations up to 200 uM result from 

hydrothermal inputs (Oremland et al., 2000). Arsenic(V) 

reducers have been found in freshwater, alkaline, and 

hypersaline environments but, thus far, only circumstantial 

evidence points to their occurrence in acidic environments 

(Oremland et al., 2001).

Colloidal Transport

Laboratory studies suggest that certain conditions within 

the subsurface may develop that enhance arsenic transport 

via mobile colloids. Increased mobility of arsenic-bearing 

colloidal material may result either from changes in the 

surface charge on colloids due to changes in subsurface 

geochemistry (e.g., Puls and Powell, 1992; Tadanier et al., 

2005) or through deflocculation and resuspension of colloidal material through dissolution of cementing agents within 

the aquifer matrix (e.g., Ryan and Gschwend, 1990; Ryan 

and Gschwend, 1992). Both processes would be facilitated 

in aquifers impacted by organic contaminants where microbial activity may be stimulated resulting in the generation of 

reducing conditions and/or the production of low molecular 

weight organic compounds that partition to fine-grained 

sediments. Due to the affinity of arsenic to the surfaces of 

iron oxyhydroxides that are commonly present as mineral 

coatings on aquifer sediments, mobilization of these colloidsized solids could potentially contribute to arsenic mobility. 

The distance of colloidal transport from an impacted zone 

is uncertain, since colloid stability may change significantly 

during transport from reduced to oxidized zones or due to 

encounter with new sediment surfaces in unimpacted zones. 

However, identification of this transport mechanism within 

an arsenic plume may play an important role relative to the 

management of the source zone or the establishment of a 

monitoring system appropriate for determining the extent 

of colloidal transport.

Site Characterization

Determining the processes controlling arsenic mobility in 

ground water and forecasting the capacity and longevity for 

attenuation is dependent on understanding the chemical 

processes controlling partitioning of arsenic onto aquifer 

solids. The aqueous and solid phase speciation of arsenic provides clues to the processes controlling solid-liquid 

partitioning, and, therefore, changes in mobility that may 

accompany chemical perturbations within the aquifer. 

Table 6.1 indicates possible natural attenuation and mobilization pathways for arsenic. 

Aqueous Measurements

Determination of arsenic chemical speciation in ground 

water provides a basis for both assessing the factors contributing to arsenic mobilization and the potential for arsenic 

partitioning to aquifer solids within and down gradient from 

the contaminant plume. The presence of arsenic as arsenite 

or thioarsenic species suggests that reductive processes 

influence arsenic mobility. In contrast, the predominance 

of the arsenate species in ground water suggests that desorption processes due to elevated pH or competition from 

other anions for sorption sites may be controlling arsenic 

mobility. Significant concentrations of methylated or other 

organo-arsenic species point to the potential influence 

of microbial processes on arsenic mobility. Thus, knowledge of aqueous arsenic speciation within the context of 

site ground-water geochemistry within and outside of the 

boundaries of the contaminant plume is important relative 

to determining the current conditions controlling arsenic 

mobility (or immobility) as well as estimating the potential 

for changes in arsenic mobility that may coincide with 

geochemical changes that occur as the aquifer returns to 

pre-contamination conditions. 
62

The total concentration of arsenic in aqueous samples can 

be assessed by a number of methods depending on the required sensitivity to meet data quality objectives (Melamed, 

2004; Melamed, 2005). Colorimetric detection based on 

the molybdenum blue method has been employed to measure inorganic arsenic for concentrations of approximately 

2 μg L-1 and higher (Johnson, 1971; Woolson et al., 1971; 

Johnson and Pilson, 1972). Determination of total arsenic 

may also be achieved using silver dithyldithiocarbamate 

as a colorimetric reagent following conversion of arsenic 

to arsine gas (Aggett and Aspell, 1976). However, these 

methods typically require conversion of all arsenic species 

in a sample to a single aqueous or gaseous species prior 

to colorimetric determination. These methods have been 

applied in the field and commercial kits are available (e.g., 

Steinmaus et al., 2006). The colorimetric methods suffer 

from analytical interferences (Stauffer, 1983; Frenzel et al., 

1994) and in some cases require the use of hazardous 

substances for test application.

More commonly, total arsenic is determined by elementspecific atomic absorption/emission or mass spectrometry 

(Eaton et al., 1998). Samples can be introduced either 

directly into the spectrometer or following conversion of 

arsenic to a hydride gas. These methods are less prone 

to analytical interferences, but complex sample matrices 

can interfere with the atomization or hydride conversion 

steps. In addition, both atomic and mass spectrometry 

may suffer from spectral interferences either from other 

elements within the sample or from molecular species that 

absorb at similar energy or possess the mass employed 

for quantification (Feldmann et al., 1999).

There are generally two approaches to analytical determination of aqueous arsenic speciation: 1) chemical 

separation of individual species in the field and subsequent determination of total arsenic in each fraction and 

2) preservation of the in-situ arsenic speciation followed by 

chemical separation and quantification of each specie in 

the laboratory. Numerous studies suggest that field separation of the various arsenic species can be achieved for 

drinking water supplies via the use of exchange resins that 

selectively pass or retain specific species (e.g., Wilkie and 

Hering, 1998; Le et al., 2000; Yalcin and Le, 2001). While 

this method is attractive, there are limitations in species 

separation (Miller et al., 2000) and potential interferences 

from competing cations or anions in contaminated water. 

This latter factor has not been sufficiently addressed for the 

complex chemistry frequently encountered in contaminant 

plumes. Thus, it is recommended that method validation 

be performed on a site-specific basis for all ground-water 

compositions that may be encountered.

The use of chemical and/or physical methods to preserve 

in-situ arsenic speciation has been investigated for a range 

of sample holding times (e.g., Haswell et al., 1985; Palacios 

et al., 1997; Eaton et al., 1998; Hall et al., 1999; Gallagher 

et al., 2004). One recurring observation from all studies is 

the importance of steps taken to prevent the precipitation of 

hydrous ferric oxide in samples collected from Fe-rich anoxic 

water in order to prevent the loss of arsenic from solution. A 

recent study provides a comprehensive review of previously 

published research along with the results of a contemporary study conducted using water samples from numerous 

sources (McCleskey et al., 2004). The analysis presented 

Table 6.1 Natural attenuation pathways for arsenic.

Attenuation Processes Mobilization Processes Characterization Approach

Precipitation of metal 

arsenates or arsenites or 

precipitation of arsenic 

sulfides

Dissolution of metal arsenates/

arsenites due to change in pH; 

dissolution of arsenic sulfides due 

to increase in pH or shift from 

reducing to oxidizing conditions.

Evaluation of arsenic speciation in ground water. 

Determination of total As in the solid matrix and 

suspected components in arsenic-bearing mineral. 

Evaluation of mineral solubility relative to groundwater chemistry and published solubility constants.

Co-precipitation of arsenic as a trace component in oxyhydroxides or 

sulfides of iron or manganese

Dissolution of host oxyhydroxide 

due to decrease in pH or shift 

from oxidizing to reducing conditions; dissolution of host sulfide 

due to shift from reducing to oxidizing conditions.

Evaluation of arsenic speciation in ground water. 

Evaluation of host precipitate formation relative to 

existing ground-water chemistry; determination of 

host mineral content in aquifer sediments via mineralogical characterization. Evaluation of arsenic 

solid phase partitioning using sequential extraction 

methodologies.

Adsorption to iron oxyhydroxides, iron sulfides, or 

other mineral surfaces 

Desorption at high pH for oxyhydroxides and sulfides; complexation/stabilization in the presence 

of DOC. Reductive dissolution of 

iron hydroxides or oxidative dissolution of iron sulfides.

Evaluation of arsenic speciation in the aqueous 

phase. Determination of total As in the solid matrix. Evaluation of arsenic solid phase partitioning 

using sequential extraction methodologies. Batch 

and column testing to determine As uptake capacity of site-specific aquifer materials with variable 

geochemical conditions.
63

in this study reconciles many of the inconsistencies from 

historical studies that examined the use of various chemical reagents for preservation of arsenic species in water. 

The authors concluded that acidification and storage of the 

sample at 4° C in a container that excludes light effectively 

preserves the As(III)/As(V) ratio for adequate periods for a 

wide range of chemical compositions of non-sulfidic ground 

water. In sulfidic waters, it appears that most of the methods 

for preservation of both total and individual arsenic species 

are inadequate either due to susceptibility of thioarsenic 

species to oxygen exposure or the precipitation of arsenic sulfides upon acidification (Smieja and Wilkin, 2003). 

However, preservation approaches have been developed 

for sulfidic waters that support reliable measurement of 

total arsenic (Smieja and Wilkin, 2003; Samanta and Clifford, 2006). Samanta and Clifford (2006) also demonstrate 

that anion exchange may be used for field separation of 

arsenite and thioarsenic species in reduced water when 

these are the dominant arsenic species. Regardless of the 

method of preservation employed, it is recommended that 

performance validation for site-specific water chemistry be 

undertaken through use of ground-water samples spiked 

with known standards in the field. The use of controlled 

laboratory solutions provides an inadequate test for the 

complexity of water chemistries that are typically encountered within contaminated ground-water systems.

Arsenic speciation in preserved aqueous samples can 

be quantified following chemical separation or masking of 

the various species. The molybedenum blue colorimetric 

method has been employed on natural samples to quantify 

As(III) and As(V) following manipulation of the oxidation 

state of As(III) (Johnson and Pilson, 1972). This method 

may be employed in the field, but is subject to interferences 

from the complex chemical matrices commonly encountered 

in contaminated ground water. More commonly, aqueous 

speciation is achieved through chromatographic separation 

of the various species followed by determination of arsenic 

by atomic absorption/fluorescence or mass spectrometry 

(Cabon and Cabon, 2000; Gomez-Ariza et al., 2000). 

Solid Phase Measurements

The implementation of an analytical approach to identify 

arsenic speciation in a soil sample is a challenging process. 

The accuracy of the analytical finding is dependent on the 

method of sample collection/preservation and the tools 

used to identify the mechanism of arsenic partitioning. It 

is recommended that the analytical protocol be designed 

to address the redox sensitivity of arsenic and the solid 

phase(s) to which arsenic may be partitioned. Tools to 

evaluate the mechanism of arsenic solid phase partitioning 

range in complexity from relatively simple chemical extractions to advanced spectroscopic techniques.

Bulk solid phase arsenic concentration can be determined 

directly on the solid sample by X-ray fluorescence spectrometry, neutron activation analysis, or following chemical digestion and analysis of arsenic in the resultant liquid phase. 

Neutron activation analysis is not commonly employed 

due to the scarcity of neutron sources required to irradiate 

the sample. X-ray fluorescence is the most attractive approach due to the relative ease of sample preparation, but 

there are potential spectral interferences. For example, the 

presence of elevated concentrations of lead can interfere 

with arsenic quantification due to overlap of the As Kα and 

the Pb Lα fluorescence peaks (Wegrzynek and Holynska, 

1993). There are approaches to correct for this interference 

provided the Pb Lβ fluorescence peak is measurable and 

free from interference. In addition, it is recommended that 

sample preparation via fusion with lithium metaborate be 

avoided due to the possible loss of arsenic through volatilization at the fusion temperature (Alvarez, 1990). There 

are various chemical digestion methods for determination of 

total arsenic (Kane, 1995; Chen and Ma, 1998), but these 

procedures do not always provide complete recovery. In 

addition, arsenic is a relatively volatile element (Yang et 

al., 1998), so sample heating to improve recovery can 

lead to losses without the appropriate precautions. Both 

approaches to quantifying total solid phase arsenic can be 

reliably applied provided the analyst is aware of the possible analytical artifacts.

There have been many applications of sequential extraction 

schemes to assess the speciation of solid phase arsenic 

(e.g., Moore et al., 1988; McLaren et al., 1998; La Force 

et al., 2000; Lumsdon et al., 2001). This approach has 

been employed primarily due to the inability to apply conventional laboratory techniques such as X-ray diffraction 

or infrared spectroscopy for detection of trace constituents 

within a soil/sediment and the limited availability of more 

advanced element-specific spectroscopies such as X-ray 

absorption spectroscopy at a synchrotron radiation facility. 

However, the lack of validated procedures and reference 

materials to test the ability of the various chemical extractants that have been employed limits the validity of selective 

extraction schemes. There are documented instances that 

clearly show selective extractants are incapable of targeting 

arsenic associated with specific solid components within a 

soil/sediment (e.g., Gruebel et al., 1988; Wilkin and Ford, 

2002). There is no single extraction protocol that can be 

recommended to chemically speciate solid phase arsenic. 

It is recommended that protocols be developed on a sitespecific basis with knowledge of the prevailing aqueous 

geochemistry and mineralogy of the ground-water aquifer 

(e.g., Ford et al., 2005).

Long-Term Stability and Capacity

The long-term stability of arsenic immobilized onto aquifer 

solids will depend on the prevailing ground-water chemistry over time relative to the conditions that existed at the 

time of immobilization. If there are significant changes 

in ground-water chemistry following immobilization, then 

the potential exists for remobilization of arsenic with the 

establishment of a new mobile plume. A general example 

of this type of situation would be if arsenic was originally 

immobilized through partitioning to mineral sulfides within 

aquifer sediments under sulfate-reducing conditions. Arsenic associated with sulfide minerals may be re-mobilized 

due to sulfide oxidation if the aquifer were to return to more 

oxidizing conditions. There are several examples of field 
64

studies that document the potential for release of arsenic 

bound to aquifer solids subsequent to changes in groundwater chemistry (Table 6.2). The examples provided in 

Table 6.2 provide general perspectives of factors that may 

increase arsenic mobility relative to the original mode for 

immobilization. While these examples are not exhaustive 

of all possible situations, they do point to the importance 

of establishing a monitoring program that is designed to 

consider the potential impact of changes in site geochemistry relative to the conditions under which arsenic was 

attenuated within the aquifer. Ultimately, the evolution of 

biogeochemical conditions in the subsurface will dictate the 

success of the MNA remedy.

Quantifying the attenuation capacity will also necessitate 

an understanding of the specific attenuation pathway(s). 

Attenuation capacity, for example, could be related to the 

extent that pH is buffered, the availability of sorptive sites in 

aquifer materials, or to the supply of electron donors needed 

to sustain microbially mediated redox conditions. For any 

proposed attenuation mechanism, there will be assumptions 

built into capacity estimations, so that uncertainty analysis 

is recommended to support capacity calculations.

Tiered Analysis

Determination of the viability of arsenic remediation in 

ground water via monitored natural attenuation will depend 

upon proper assessment of contaminant loading to the 

aquifer and prevailing geochemistry and mineralogy within 

the contaminant plume and the down gradient zone prior 

to the point(s) of compliance. The goal of site assessment 

is to demonstrate the process(es) controlling arsenic sequestration onto aquifer solids and the long-term stability 

of solid phase arsenic as a function of existing and anticipated ground-water chemistry. A recent technical review 

highlights limited instances where natural attenuation was 

chosen as a component of the remedy for cleanup of arsenic 

in ground water (Reisinger et al., 2005). However, as noted 

in the review, site characterization conducted to support 

selection of natural attenuation as a remedy at these sites 

did not fully identify the site-specific immobilization process 

and the long-term stability of the sequestered arsenic. The 

following tiered analysis structure for site characterization 

provides a technically defensible approach to evaluate 

candidate sites and define the potential limitations of MNA 

as part of a remedy for ground-water cleanup.

Tier I - Site characterization under Tier I will involve demonstration that the plume is static or shrinking, has not reached 

compliance boundaries, and does not impact existing water 

supplies. Once this is established through ground-water 

characterization, evidence is collected to demonstrate As 

partitioning to aquifer solids within the plume. If natural 

attenuation processes are active throughout the plume, it 

is anticipated that this would be reflected in an observed 

increase in solid phase concentrations within regions 

of the plume with higher aqueous concentrations, e.g., 

near the source term. This field partitioning data may be 

supplemented by geochemical modeling that incorporates 

measured water chemistry (e.g., pH, Eh, and major ion 

chemistry) throughout the plume to assess the potential for 

solubility control by metal arsenates/arsenites or arsenic 

sulfides. Since identification of the chemical speciation of 

arsenic in water samples and aquifer sediments is critical 

towards determining the attenuation mechanism(s), it is 

recommended that precautions be taken to preserve arsenic 

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

Location and Setting Original Conditions

(Arsenic Immobile)

Altered Conditions

(Arsenic Mobilized)

Unconsolidated aquifer consisting 

of glacial outwash, glaciomarine 

clay, and till overlying bedrock; 

New Hampshire a

Arsenic bound to poorly crystalline iron hydroxides in glaciomarine clay under oxidizing conditions (Eh ~ 400 mV)Reductive dissolution of iron hydroxides due 

to microbially driven Fe-reducing conditions stimulated by organic carbon in landfill 

leachate

Suwannee Limestone, Upper 

Floridan consolidated aquifer; 

Florida b

Arsenic hosted in pyrite that is 

most abundant in high porosity 

zones; anoxic aquifer 

Oxidative dissolution of pyrite due to injection of oxygenated water

Unconsolidated, glacial outwash aquifer consisting of 

coarse(quartz) sand and gravel 

with Fe and Al oxide and/or 

silicate mineral coatings; New 

Hampshire c

Arsenic adsorbed to mineral 

coatings under oxidizing conditions

Adsorbed arsenic mobilized due to desorption in the presence of elevated phosphate 

derived from sewage disposal via land application coupled with reductive dissolution 

of Fe-bearing mineral coatings

a Delemeos et al., 2006; b

 Price and Pichler, 2006; c Kent and Fox, 2004
65

speciation during collection, preservation, and processing of 

collected samples (See recommendations and application 

of these methods in Wilkin et al., 2002, Ford et al., 2005, 

and Wilkin, 2006). Identification of active sequestration to 

prevent arsenic migration in ground-water provides justification for proceeding to Tier II characterization efforts. 

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

of attenuation are determined. Estimates of a site attenuation rate(s) can be assessed via a well transect along 

the ground-water flow path. In addition, time-series data 

may be collected at one or more monitoring points within 

the plume. This information will allow assessment of the 

relative timescales for contaminant immobilization and fluid 

transport and determination of whether remediation objectives can be met within the required regulatory time frame. 

In addition, the mechanism(s) for attenuation need to be 

identified under this stage of site characterization. This 

effort may require determination of the chemical speciation 

of aqueous and solid phase As, which may be approached 

according to the following scheme:

1) Determination of arsenic solution speciation 

via direct analytical measurements in combination with speciation calculations based on 

characterized ground-water chemistry;

2) Calculation of the saturation state of ground 

water relative to measured aqueous chemistry complimented by the possible isolation 

of discrete As mineral phases via density 

separations (or other schemes) in regions of 

the aquifer with highest solid phase concentrations;

3) Determination of aquifer mineralogy to determine the relative abundance of components 

with documented capacity for As sorption (e.g., 

Amonette, 2002);

4) Identification of arsenic association(s) with 

the various solid phase components of aquifer 

solids through combination of chemical extractions with microscopic/spectroscopic confirmation of phase associations, and;

5) Demonstration of concurrence between the site 

conceptual model and mathematical model(s) 

that describe arsenic removal mechanism(s).

It is recommended that identification of arsenic chemical 

speciation in aqueous and solid matrices be conducted using samples collected in a manner that preserves the in-situ 

mineralogy and speciation of arsenic. The demonstration 

of concurrence between conceptual and mathematical 

models describing arsenic transport will entail development 

of site-specific parameterization of the chemical processes 

controlling arsenic solid phase partitioning.

Tier III - Once the partitioning mechanism(s) have been 

identified for the site, the subsequent characterization effort 

under Tier III will involve determination of the stability of 

immobilized As and the capacity of the aquifer to sustain 

continued uptake. It is recommended that the stability of 

immobilized As be tested based on the anticipated evolution 

of ground-water chemistry concurrent with plume shrinkage. For example, changes in ground-water redox potential 

due to oxygen intrusion or the occurrence of anaerobic 

microbial processes can exert a significant influence on As 

partitioning to iron sulfides or iron oxyhydroxides, respectively. Therefore, it is recommended that sediment leach 

tests be conducted to characterize the magnitude of As 

mobilization as a function of ground-water redox chemistry 

representative of existing and anticipated site conditions. 

This may involve the use of microcosm tests that stimulate 

in-situ microbial populations toward the development of 

redox conditions considered deleterious for continued As 

immobilization. It is recommended that the capacity for 

As uptake onto aquifer solids be determined relative to the 

specific mechanism(s) identified in Tier II. For example, if 

site characterization under Tier II indicated that co-precipitation of As with iron sulfide due to microbial degradation of 

organic compounds coupled with sulfate reduction occurs 

within the aquifer, then it is recommended that the mass 

distribution of organic carbon, sulfate and ferrous iron to 

support this reaction within the aquifer be determined. This 

site-specific capacity can then be compared to As mass 

loading within the plume in order to assess the longevity 

of the natural attenuation process. If site-specific tests 

demonstrate the stability of immobilized As and sufficient 

capacity within the aquifer to sustain As attenuation, then 

the site characterization effort can progress to Tier IV. For 

cases where contaminant stability is sufficient but aquifer 

capacity is insufficient for capture of the entire plume, then 

a determination of the benefits of contaminant source reduction may be necessary.

Tier IV - Finally, under Tier IV a monitoring plan is established along with contingency plans in the event of MNA 

failure. It is recommended that the monitoring plan be 

designed to establish both continued plume stability and to 

identify changes in ground-water chemistry that may lead 

to re-mobilization of attenuated As. The specific chemical 

parameters to be monitored will include those identified 

under Tier III that may halt As partitioning to aquifer sediments and/or result in solubilization of either discrete As 

precipitates or aquifer minerals that sequester As from 

ground water. For example, solution phase parameters 

that could alter either As precipitation or adsorption include 

increases in the concentration of competing anions, such 

as phosphate, in combination with changes in ground-water pH. In contrast, the concentration of dissolved iron or 

sulfate may indicate the dissolution of an important sorptive 

phase within the aquifer (e.g., reductive dissolution of iron 

oxyhydroxides or oxidative dissolution of sulfides). Changes 

in these parameters may occur prior to observed changes 

in solution As and, thus, serve as monitoring triggers for 

potential MNA failure. In this instance, a contingency plan 

can be implemented that incorporates interventive strategies to arrest possible plume expansion beyond compliance 

boundaries. Possible strategies to prevent plume expansion 

include pump and treat operations, installation of reactive 

barriers to enhance uptake capacity perpendicular to the 

direction of plume advance, or enhancement of natural attenuation processes within the aquifer through the injection 

of soluble reactive components.
66

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71

Selenium

Chunming Su, Robert G. Ford, Richard T. Wilkin

Occurrence and Distribution

Selenium contamination can be derived from weathering of 

natural deposits as well as discharge from mining, smelting, and coal/oil combustion. Most soils contain between 

0.1 to 2 mg kg-1, but soils originating from the Upper Cretaceous marine sedimentary rocks (shale) show elevated 

concentrations of selenium in about 80,000 km2 of land 

in the 17 western states of the United States (Nolan and 

Clark, 1997). In California, shale materials in the Coastal 

Ranges along the entire western margin of the San Joaquin 

Valley contain excessive and varied amounts of selenium 

with median values as high as 6.5 to 8.7 mg kg-1 (Presser 

and Barnes, 1984). Complex interactions among irrigated 

agricultural practices and physical and chemical processes 

have contributed to high selenium concentrations (20 to 

1000 μg L-1) in ground water underlying low-altitude agricultural areas of the western San Joaquin Valley (Deverel 

et al., 1984; Seiler, 1997). 

Incidences of waterfowl mortality and reproductive failure 

at Kesterson Reservoir, San Joaquin Valley, California was 

discovered in 1983 and was attributed to elevated selenium 

concentrations in irrigation drainage (Ohlendorf et al., 

1986). More follow-up studies have revealed that elevated 

selenium concentrations in water from irrigation projects 

in the Western USA are largely caused by application of 

irrigation water to soils derived from shales of Cretaceous 

age (Naftz, 1996). 

There are 21 former uranium mill sites designated under 

the Uranium Mill Tailings Radiation Control Act (Federal 

Register, 1995) that show ground-water contamination with 

toxic elements including selenium, arsenic, manganese, 

molybdenum, uranium, vanadium, and zinc. Contamination 

of groundwater with selenium also occurs from surface 

coal mining in Wyoming (Blaylock et al., 1995). Waste rock 

from phosphorite mining in southeastern Idaho contains 

selenium, cadmium, vanadium, and other metals. Selenium 

concentrations in the hundreds of parts per million have 

been found in soils, plants, and seeps in close proximity 

to some mine waste dumps. Mortality of livestock due to 

selenosis poisoning has been linked to the contamination of 

water and vegetation by mine waste (Piper et al., 2000). 

Plume Characteristics

The general patterns in subsurface selenium transport that 

may be anticipated under differing geochemical conditions 

are represented by a series of field investigations that have 

been conducted in aquifers with redox chemistry grading 

from oxic to strongly reducing. For oxic conditions (8 mg L-1

O2, pH ~5, no detectable Fe2+ or sulfide), the field study 

conducted by Kent et al. (1995) demonstrated that selenium 

(as selenate) transport will likely be dominated by adsorption reactions with aquifer sediment. While adsorption will 

tend to retard selenium transport, this process will not likely 

prevent plume expansion. For mildly reducing conditions 

(0.032 mg L-1

 O2, pH ~6, 17-28 mg L-1 Fe2+, ≤0.034 mg L-1

sulfide), Kent et al. (1994) presented results that indicated 

little selenate attenuation due to insufficient conditions to 

support either biotic or abiotic reduction of selenate to less 

mobile forms. In contrast to selenium transport behavior 

within the oxic zone of this aquifer, there was little retardation of selenium transport relative to a conservative tracer 

(bromide). In the absence of conditions insufficient for selenium reduction to Se(0) or more reduced forms, the higher 

pH and greater concentrations of anions that compete for 

adsorption sites on aquifer minerals resulted in conservative transport of selenate within the suboxic portion of the 

aquifer. For highly reducing conditions (≤0.006 mg L-1

 O2, 

pH ~7, ≤25 mg L-1 Fe2+, ≤0.5 mg L-1 sulfide), field studies 

conducted in the shallow aquifer underlying a wetland system at the Kesterson Reservoir, California demonstrated 

significant attenuation of selenium (White et al., 1990; 

Benson et al., 1991; White et al., 1991). This attenuation 

apparently resulted from reduction of selenate and/or 

selenite to Se(0) or more reduced forms upon interaction 

with biotic/abiotic components of the aquifer sediments in 

the reducing ground water. These observations generally 

indicate that the attenuation of selenium will be correlated 

with the extent of its reduction within an aquifer. 

Remedial Technologies 

Active removal of selenium from ground water may be 

achieved either through above ground or subsurface (insitur) treatment processes. Examples of these approaches 

include 1) above-ground treatment operations where selenium is removed via adsorption, ion exchange, or chemical 

reduction, 2) phytoremedation for shallow ground-water 

systems either through crop planting or the construction 

of wetland systems that intercept ground-water discharge, 

and 3) installation of permeable reactive barriers in the 

subsurface to intercept a migrating plume.

For above ground treatment operations, layered double 

hydroxide precipitates have been used as sorbents for removing both selenate and selenite from water (O’Neil et al., 
72

1990). Magnesium-aluminum and zinc-aluminum layered 

double hydroxides show high affinity for both selenate and 

selenite (You et al., 2001) with uptake of these anions via 

exchange with structural interlayer anions. Chemical reduction, primarily through stimulation of microbial processes, 

has been employed for selenium removal from contaminated 

water. An algal bacterial selenium removal facility was designed and constructed at the Panoche Drainage District in 

California during 1995-1996 to remove selenium and nitrate 

from drainage water (Green et al., 2003). Mass removals of 

total soluble selenium of 77% or greater were achieved over 

a three-year period. Selenate was removed by assimiliatory 

and dissimiliatory bacterial reduction. Laboratory tests have 

shown the benefit of using rice straw to promote selenate 

removal from drainage water through microbial reduction 

of selenate to selenite and subsequently to colloidal Se(0) 

(Zhang and Frankenberger, 2003). In another study, a 

flow-through wetland system was constructed consisting 

of ten unlined cells that were flooded continuously. The 

global mass balance showed that on average 59% of the 

total inflow selenium was retained within the wetland cells, 

whereas selenium outputs included outflow (35%), seepage (4%), and volatilization (2%) (Gao et al., 2003a). The 

major selenium removal mechanism was the reduction of 

selenate to Se(0) and immobilization into the organic phase 

of the sediments (Gao et al., 2003b). However, it has been 

noted that in some wetland systems, biological volatilization of selenium may account for 10-30% of the selenium 

removed from contaminated water (Allen, 1991; Hansen et 

al., 1998; Zhang and Frankenburger, 1999).

The use of phytoremediation as a treatment technology 

involves removal of selenium from contaminated water 

either through absorption into the plant mass or volatilization following uptake and conversion to organic species 

(Banuelos et al., 2002). Potential crops tested for the 

phytoextraction of selenium in central California include 

two moderate selenium accumulators, Indian mustard and 

canola (Banuelos et al., 1993). Selenium hyperaccumulators have been identified, e.g., Astragalus bisulcatus, which 

can accumulate up to 0.65% (w/w) selenium. Pickering et 

al. (2003) have identified high concentrations of the nonprotein amino acid Se-methylseleno-cystein in young leaves of 

this hyperaccumulator, but in more mature leaves, selenate 

predominates. Another selenium hyperaccumulator, Stanleya pinnata (Brassicaceae) is a perennial that responded 

favorably to repeated cutting in the greenhouse, a trait 

that could prove valuable in field-scale phytoremediation 

(Parker et al., 2003). 

The use of zerovalent iron [Fe(0)] for removal of selenium in 

aqueous systems has been demonstrated in both laboratory 

and field applications. Earlier studies show that selenate 

can be reduced to a lower oxidation state by metallic iron in 

water (Baldwin et al., 1981). Murphy (1988) reported that 

selenate may be removed by Fe(0) through the formation 

of Se(0). Selenium in the form of selenocyanate (SeCN-

) 

in oil refinery wastewater and artificial wastewater could be 

removed using Fe(0) through the formation of Se(0) and 

ferrous selenide (FeSe). The possible chemical reaction 

between SeCN- and Fe(0) included deselenation of SeCNand electrochemical reduction of Se(0) to Se(-II) (Meng 

et al., 2002). Morrison et al. (2002) have demonstrated 

selenium removal from ground water using Fe(0) in a field 

installation at a uranium mill tailings repository. Three 

treatment cells [Cercona foamed Fe(0) plates, steel wool, 

and Peerless granular Fe(0)] were operated for passive 

removal of selenium along with arsenic, manganese, molybdenum, uranium, vanadium, and zinc from ground water 

(Morrison et al., 2002). Significant removal was achieved 

in the treatment cell using Cercona foamed Fe(0) plates 

(reduction from 202 μg L-1 to 6 μg L-1 dissolved Se), and 

the cell performed for more than three years. 

Regulatory Aspects

Selenium was ranked number 144 out of a total of 275 on 

the CERCLA Priority List of Hazardous Substances in 2001, 

which was prepared by the Agency for Toxic Substances 

and Disease Registry, Centers for Disease Control and Prevention. The USEPA has set the maximum contaminant level 

(MCL) at 0.05 mg L-1 for drinking water (USEPA, 2006a; http: 

//www.epa.gov/waterscience/criteria/drinking/dwstandards.

pdf). The USEPA has found selenium to potentially cause 

the following health effects when people are exposed to it 

at levels above the MCL: hair or fingernail loss, numbness 

in fingers or toes, nervous and circulatory systems problems, and kidney and liver damage. For non-potable water 

sources, ambient water quality criteria (AWQC) that are 

protective of aquatic life may serve as alternative cleanup 

goals. For selenium, current statutes list the chronic criterion for fresh waters as 0.005 mg L-1

 with recommendation 

for an acute criterion pending review (USEPA, 2006b; http: 

//www.epa.gov/waterscience/criteria/nrwqc-2006.pdf). An 

example of where this criterion may apply is a site where 

contaminated ground water discharges to surface water.

Selenium (Se) is an essential nutrient for humans and 

vertebrates, but is only required in small amounts and 

has a very narrow range between deficient and toxic 

levels (National Research Council, 1983). It is found 

in the enzyme glutathione peroxidase that inhibits the 

oxidative role of peroxides and hydroperoxides, thereby 

protecting immunocomponent cells and slowing down 

aging processes. Selenium also plays an important role 

in anticarcinogenic activity and prevention of heavy metal 

toxic effects. Selenium detoxifies mercury in humans via 

possible formation of 1: 1 Hg-Se compounds (Kosta et al., 

1975). Selenium yeast can both prevent the accumulation of 

arsenic in the human body and rectify the damages (Wang 

et al., 2001). Selenium supports efficient thyroid hormone 

synthesis and metabolism and protects the thyroid gland 

from damage by excessive iodide exposure (Zimmermann 

and Kohrle, 2002). Epidemiological studies have indicated 

an inverse relationship between selenium intake and the 

incidence of certain cancers; blood or plasma levels of 

selenium are usually lower in patients with cancer than 

those without this disorder (Whanger, 2004). An endemic 

human disease referred to as Keshan disease in certain 

regions of China was caused by selenium deficiency (Chen 

et al., 1980). Keshan disease is a juvenile cardiomyopathy 

that presents as congestive heart failure in infants and 
73

young children. This disease has been virtually eliminated 

by supplying sodium selenite pills to those at risk. Another 

selenium deficiency disease has been also reported in 

areas of China and referred to as Kaschin-Beck disease, 

which is an osteoarthropathy, a generative articular disease 

caused by oxidative damage to cartilage that leads to 

deformation of bone structure (Ge and Yang, 1993). 

In contrast, excess selenium is very toxic and can cause 

selenium poisoning (selenosis) in humans and animals. 

Consumption of feed containing greater than 5 mg of Se 

kg-1 leads to selenium poisoning in animals (Anderson et 

al., 1961; National Research Council, 1980). Selenium 

toxicity disorders in livestock known as “alkali disease” 

and “blind staggers” were widely recognized in the 1930s 

in South Dakota in areas of high soil selenium (Magg and 

Glen, 1967). Sheep grazing in areas of phosphate mining 

operations in southeastern Idaho have died as a result of 

high selenium concentrations in forage and water (Fessler 

et al., 2003). Selenium toxicity also has been reported in 

some regions of Australia as a result of livestock feeding 

on selenium accumulative plant species that contain selenocystathionine (Tinggi, 2003). Endemic selenium poisoning of humans characterized by loss of nails and hair was 

reported in Yutangba of central China in the 1960s, which 

was associated with the use of carbonaceous shales (stone 

coal) high in selenium in the form of elemental selenium 

(Zhu et al., 2004). 

Geochemistry and Attenuation Processes

Aqueous Speciation

Selenium is a metalloid exhibiting physical and chemical 

properties between that of metals and nonmetals. It chemically resembles sulfur and exists in organic and inorganic 

chemical forms. Inorganic species include selenide [Se(-II)], 

elemental selenium [Se(0)], selenite [Se(IV)], and selenate 

[Se(VI)]. Organic species include methylated compounds, 

selenoamino acids, selenoproteins and their derivatives. 

The speciation of selenium is greatly influenced by the pH 

and redox conditions of the environment (Figure 7.1). For 

example, Se(-II) exists in a reducing environment as hydrogen selenide (H2

Se) and as metal selenides. Reduction of 

selenate to selenite and Se(0) has been shown to decrease 

its mobility in saline, mildly alkaline groundwater (White 

et al., 1991). When dissolved in water, H2

Se can oxidize 

to elemental selenium. Elemental selenium is stable in a 

reduced environment, but it can be oxidized to selenite and 

to selenate by a variety of microorganisms (Sarathchandra 

and Watkinson, 1981). 

Various strains of bacteria have been identified to facilitate 

selenate reduction in soil and sediment systems. Two 

microbial processes, namely methylation of selenium and 

reduction of both selenate and selenite to Se(0), have a 

major influence on the fate and mobility of this element in the 

environment (Dungan and Frankenberger, 1999; Dungan 

et al., 2003). Methylation of selenium, and subsequent 

selenium volatilization, leads to dissipation of soil selenium 

to the atmosphere. Environmental factors such as the 

existing microbial community, pH, temperature, moisture, 

and organic amendments control the rate of selenium 

volatilization from seleniferous soils (Frankenberger and 

Karlson, 1989; Zhang and Frankenburger, 1999). Under 

flooding conditions, part of the methylated selenium may be 

transported in water, thus decreasing selenium volatilization 

to the atmosphere (Zhang and Frankenburger, 1999). 

The addition of organic amendments to soils has been 

reported to stimulate indigenous microbes to methylate 

selenium (Abu-Erreish et al., 1968; Frankenberger and 

Karlson, 1989); whereas, organic substrates added to 

ponded sediments have been found to accelerate the 

reduction of selenate and selenite to Se0

 (Tokunaga et al., 

1996) and similar effects were reported in laboratory batch 

experiments (Zhang et al., 2003). 

Figure 7.1 Eh-pH diagram for selenium at 25 °C using 

thermodynamic data from Séby et al. (2001). 

ΣSe=10-5 (790 μg L-1). Solid stability field for 

elemental selenium is shaded pink.

Mechanisms for selenium reduction by microbes are complex as it occurs under both aerobic (Lortie et al., 1992) 

and anaerobic conditions (Oremland et al., 1989; Oremland 

et al., 1990; Tomei et al., 1992). Both dissimilatory and 

detoxification mechanisms are possible (Oremland, 1994). 

The occurrence of sequential reduction of selenate to selenite and then to Se(0) is suggested after amendment of 

contaminated soils with barley straw under field capacity 

moisture conditions (Camps Arbestain, 1998). The rates 

of selenate to selenite reduction in waste waters is proportional to their respective concentrations in solution and 

also to the amount of the microbial biomass (Rege et al., 

1999). A recent study has documented the occurrence of 

both intracellular and extracellular Se(0) granules in three 

phylogenetically and physiologically distinct bacteria that 

are able to respire selenium oxyanions, suggesting that 

this phenomenon appears to be widespread among such 

bacteria (Oremland et al., 2004). The metal sites of selenate 

reductase from Thauera selenatis have been characterized 
74

(Maher et al., 2004); the enzyme was found to contain Se in 

a reduced form (probably organic) and the Se is coordinated 

to both a metal (probably Fe) and carbon. Assessment of 

Se(IV) and Se(0) reduction in anaerobic estuarine sediment microcosms demonstrated the sequential formation of 

Se(0) and ultimately Se(-II) during incubation with elemental 

selenium and/or lactate as electron donors for microbial 

reduction (Herbel et al., 2003). Dissolved Se(-II) did not 

accumulate in sediment pore water during incubation due 

precipitation with ferrous iron to form FeSe(s), which was 

determined via solid phase characterization using X-ray 

absorption spectroscopy.

Solubility

Selenium has the potential of forming precipitates for all 

of its oxidation states (Seby et al., 2001; see Figure 7.2). 

For selenate and selenite, this includes precipitates with 

common major cations in ground water (Ca, Mg) as well 

as transition metals (Fe, Mn) or heavy metals that may be 

anticipated in contaminant plumes (Rai et al., 1995; Sharmasarkar et al., 1996). These phases are anticipated to 

primarily be significant in situations where selenium concentrations are highly elevated. Reduction to form elemental 

selenium [Se(0)] can result in very low concentrations of 

dissolved selenium. In general, it has been observed that 

selenium reduction to insoluble Se(0) results in immobilization and stabilization of this element in the soil matrix, 

since the re-oxidation reaction of Se(0) to soluble selenate 

and selenite is relatively slow (Tokunaga et al., 1994; Zawislanski and Zavarin, 1996). Abiotic reduction of selenite 

to Se(0) was also suggested in sulfate-reducing bacteria 

(SRB) biofilms (Hockin and Gadd, 2003). Elemental selenium and elemental sulfur were found to precipitate outside 

SRB cells. Further reduction to selenide [Se(-II)] can lead 

to precipitation of metal selenides, including ferrous iron 

and manganese selenides, similar to the formation of metal 

sulfides under sulfate-reducing conditions.

While not commonly observed, it is anticipated that many 

suboxic geological environments contain green rust, which 

is a mixed ferrous, ferric hydroxide that also contains 

interlayer anions such as sulfate and carbonate in its 

structure (Feder et al., 2005). Identification of green rusts 

in sediments and soils is hampered by the rapid oxidation 

of green rusts by atmosphere oxygen, and for this reason, 

they have not been commonly reported. However, recent 

thermodynamic and spectroscopic studies give direct evidence for the existence of green rusts in soils (Hansen et 

al., 1994; Trolard et al., 1997; Feder et al., 2005). Due to 

high reactivity, green rust minerals are envisioned as potential reducing agents of a number of contaminants such 

as nitrate, chromate, and selenate. Direct evidence for the 

formation of reduced selenium species in anoxic sediments 

via abiotic redox reactions with sulfate green rust was 

provided using X-ray absorption near-edge spectroscopy 

(XANES) and Fourier-transform extended X-ray absorption 

fine structure (EXAFS) spectroscopy (Myneni et al., 1997). 

The mechanism of selenate reduction was described by 

the following equation: 

 HSeO4

-

 + 4Fe4

IIFe2III(OH)12SO4•3H2

O →

 HSe- + 8Fe3O4 + 4SO4

2- + 8H+ + 32H2

O (2)

in which sulfate green rust was oxidized to form magnetite, 

whereas selenate was reduced to Se(0) and subsequently 

to selenide (herein referred to as the ‘adsorption-reduction’ 

pathway). In addition, a laboratory study has demonstrated 

that a significant fraction of dissolved selenate can be coprecipitated with Fe(II) and Fe(III) ions to form Fe(II)-Fe(III) 

hydroxyselenate green rust with simultaneous reduction of 

an equal amount of selenate anions to selenite anions (Refait et al., 2000). In the subsurface environment, selenate 

reduction by coprecipitation and adsorption pathways can 

occur when selenium-contaminated sediment becomes 

reducing (Pickering et al., 1995). In the coprecipitationreduction pathway, reductive dissolution of Fe(III) oxides 

precipitates green rust with selenate followed by selenate 

reduction to Se(0) and selenide. Reduction of selenium 

oxyanions to Se(0) has also been observed in the presence of iron sulfides (Bruggeman et al., 2005) and ferrous 

hydroxide (Zingaro et al., 1997). 

Figure 7.2 Eh-pH diagram for selenium at 25 °C using 

thermodynamic data from Séby et al. (2001). 

System Se-H2

O-Fe-Ca, with ΣSe=10-5 M 

(790 μg L-1), ΣFe=10-4 M (5.6 mg L-1), and 

ΣCa=10-2 M (400 mg L-1). Solid stability fields for elemental selenium, hydrous 

calcium selenite, and ferrous selenide are 

shaded pink. FeSe2 was suppressed (data 

not available in the Séby et al. review); 

however, the stability field of the diselenide 

would be intermediate between elemental 

selenium and FeSe.

Adsorption

A comprehensive summary of selenate and selenite adsorption behavior on individual soil minerals (Fe, Al, and Mn 

oxides, kaolinite, and calcite) and whole soils throughout the 

United States is documented in Zachara et al. (1994). This 
75

research demonstrated the importance of Fe and Al oxide 

surfaces for selenium adsorption onto aquifer sediments 

and highlighted the dependence of the extent of adsorption on the pH of ground water and the presence of anions 

that compete for adsorption sites. Additional review of the 

published literature is provided below.

Selenate has been shown to behave like sulfate with minimal adsorption and high mobility (Goldberg and Glaubig, 

1988; Neal and Sposito, 1989); whereas, selenite behaves 

analogously to phosphate, with greater adsorption than 

selenate (Neal et al., 1987; Barrow and Whelan, 1989a; 

Zhang and Sparks, 1990). Adsorption of selenite on goethite 

decreases with increasing pH, with decreasing selenite concentration, and with competing anions such as phosphate, 

silicate, citrate, molybdate, carbonate, oxalate, and fluoride 

(Balistrieri and Chao, 1987). More selenite is adsorbed onto 

montmorillonite than on kaolinite (Frost and Griffin, 1977). 

Selenite adsorption in seleniferous soils are decreased in 

the presence of sulfate, nitrate, and phosphate (Pareek 

et al, 2000). Selenite sorption by aluminum hydroxides is 

adversely affected by organic acids (Dynes and Huang, 

1997). Selenite selectively adsorbs at the carbonate (CO3

2-) 

site on calcite (CaCO3) via ionic exchange, forming a twodimensional solid-solution of the form Ca(SeO3

)

x

(CO3)1-x at 

the interface; under identical chemical conditions, selenate 

adsorption is inhibited (Cheng et al., 1997). An earlier 

study showed selenate substitution in calcite also occurred 

(Redder et al., 1994). Desorption of selenate is faster and 

more nearly complete than selenite after adsorption and 

incubation in soil (Barrow and Whelan, 1989b).

Mechanisms of selenium adsorption have been studied from 

both macroscale batch and microscale spectroscopic approaches. The presence of either selenate or selenite lowers 

the electrophoretic mobility and decreases the point of zero 

charge of amorphous iron hydroxide (am-Fe(OH)3

)and goethite (α-FeOOH), suggesting inner-sphere complexation for 

both selenate and selenite species (Su and Suarez, 2000). 

Both in situ attenuated total reflection – Fourier transform 

infrared (ATR-FTIR) and diffuse reflectance infrared Fourier 

transform (DRIFT) spectra show bidentate complexes of 

selenate with am-Fe(OH)3 and the DRIFT spectra of selenite 

on goethite show bridging bidentate complex of selenite. 

These results are consistent with an earlier in situ extended 

x-ray absorption fine structure (EXAFS) spectroscopic 

study (Manceau and Charlet, 1994) that shows selenate 

forms an inner-sphere binuclear bridging surface complex 

on hydrous ferric oxide and goethite. On the contrary, an 

earlier EXAFS study (Hayes et al., 1987) concluded that 

selenate forms an outer-sphere surface complex on goethite. A recent combined data set of Raman and ATR-FTIR 

spectra indicate that both inner- and outer-sphere surface 

complexes of selenate occur on goethite, as predominantly 

monodentate inner-sphere surface complexes at pH < 6, 

and as predominantly outer-sphere surface complexes at 

pH > 6 (Wijnja and Schulthess, 2000). 

Site Characterization

The conditions that favor the mobility of selenium in the 

environment with respect to adsorption are alkaline pH, 

high selenium concentrations, oxidizing conditions, and high 

concentrations of additional anions that strongly adsorb. 

Most of the reductive capacity in aquifers resides in the 

sediments (Barcelona and Holm, 1991). It is thus critical 

to assess the predominant ground-water chemistry across 

the plume as well as the biogeochemical characteristics of 

the aquifer sediments that may lead to selenium attenuation. Table 7.1 provides a summary of potential attenuation 

processes that may be active within an aquifer along with 

a general approach to site characterization to identify the 

active attenuation process(es).

Aqueous Measurements

The mobility of selenium depends mainly on the different 

chemical forms in which it is present. Information on the 

chemical speciation of selenium in ground water provides 

part of the context for understanding the processes that may 

control its attenuation within the aquifer. Measurements that 

can be used to assess the chemical speciation of selenium 

include determination of its total dissolved concentration 

along with quantification of individual chemical species 

(inorganic or organic) that may occur within the plume. 

Published USEPA methods that may be employed for the 

determination of the total dissolved concentration of selenium are documented in Table 7.2. Of these methods, only 

the gaseous hydride technique can differentiate between 

inorganic chemical selenium species, namely selenate 

[Se(VI)] and selenite [Se(IV)], since only selenite forms 

a gaseous hydride that is detected using this technique. 

In this case, selenate may be inferred by the difference 

between measured selenite and total dissolved selenium. 

However, as previously reviewed, there are other inorganic 

(e.g., HSe-) and organoselenium species that may occur 

within ground water. Thus, additional speciation methods 

may be needed in order to properly characterize the distribution of dissolved selenium species. 

There are several documented approaches to determine 

the chemical speciation of selenium in water. These approaches can be classified into two general categories: 

1) those that use a chemical/mass specific detector following chromatographic separation of individual aqueous 

species (hyphenated techniques), and 2) those that provide 

direct and indirect detection of individual species prior to or 

following chemical conversion of individual species within 

the sample (e.g., voltammetric techniques). Recent comprehensive reviews are available for hyphenated techniques 

that employ atomic absorption/fluorescence spectrometry 

(Capelo et al., 2006) or inductively coupled plasma-mass 

spectrometry (B’Hymer and Caruso, 2006). For the latter 

technique, Tirez et al. (2000) provide a useful overview of 

possible isobaric (similar ion mass) and polyatomic interferences (due to argon bonding to sample matrix elements 

within plasma) that can bias analytical results for individual 

chemical species of selenium. There are several published 

procedures that employ voltammetry for the detection of 

selenium in natural water samples (Locatelli and Torsi, 2000; 

Ochsenkuhn-Petropoulou and Tsopelas, 2002; Bertolina 

et al., 2006). As outlined by Ochsenkuhn-Petropoulou 

and Tsopelas (2002), not all aqueous selenium species 
76

Table 7.1 Natural attenuation and mobilization pathways for selenium.

Attenuation Processes Mobilization Processes Characterization Approach

Biotic (microbial) or abiotic 

reduction by reduced Febearing minerals of selenate/selenite to elemental 

Se(0) and/or selenide with 

precipitation as a metal 

selenide.

Cessation of microbial processes via 

changes in nutrient supply within the 

ground-water flow path; reduction 

in the mass of reduced Fe-bearing 

minerals within the aquifer due to 

changes in redox conditions or supply 

of soluble iron. Oxidation of Se(-II) 

or Se(0) upon exposure to oxygen or 

other oxidants.

Determination of water chemistry, microbial 

populations, and/or sediment mineralogy 

linked to selenium reduction along with 

spatial and temporal variability of required 

chemical/microbial components relative to 

selenium transport pathway(s). Determination of total concentration and selenium 

chemical speciation in aquifer sediments.

Precipitation of metal selenates or selenitesDissolution of metal selenates/selenites due to increased pHEvaluation of selenium speciation in ground 

water. Determination of total Se in the solid 

matrix and suspected components in arsenic-bearing mineral. Evaluation of mineral 

solubility relative to ground-water chemistry 

and published solubility constants.

Adsorption to iron oxyhydroxides, iron sulfides, or other 

mineral surfaces

Desorption at high pH for oxyhydroxides 

and sulfides; adsorption inhibition in the 

presence of DOC or competing anions. 

Reductive dissolution of iron hydroxides 

or oxidative dissolution of iron sulfides.

Evaluation of selenium speciation in the aqueous phase. Determination of total Se in the 

solid matrix. Evaluation of selenium solid 

phase partitioning using single or sequential 

extraction methodologies. Batch and column 

testing to determine Se uptake capacity of 

site-specific aquifer materials with variable 

geochemical conditions.

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

Method Name and Number Source

Inductively Coupled Plasma-Atomic Emission 

Spectrometry, 6010B, Revision 2

December 1996 

http://www.epa.gov/epaoswer/hazwaste/test/pdfs/6010b.pdf

Inductively Coupled Plasma-Mass Spectrometry, 

6020, Revision 0

September 1994 

http://www.epa.gov/epaoswer/hazwaste/test/pdfs/6020.pdf

Selenium (Atomic Absorption, Furnace Technique), 7740, Revision 0 September 1986 a

Selenium (Atomic Absorption, Gaseous Hydride), 

7741a, Revision 1

September 1994 

http://www.epa.gov/epaoswer/hazwaste/test/pdfs/7741a.pdf

Selenium (Atomic Absorption, Borohydride Reduction), 7742, Revision 0September 1994 

http://www.epa.gov/epaoswer/hazwaste/test/pdfs/7742.pdf

a Available electronically at http://web1.er.usgs.gov/nemi/method_summary.jsp?param_method_id=5239.
77

are electrochemically active [e.g., selenate and (CH3)

2

Se]. 

Thus, this analytical approach requires the development and 

testing of sample preparation schemes that can separate 

individual electrochemically active and inactive species 

and convert the latter to active forms. Given this analytical 

limitation, a potential advantage of this approach is that 

the instrumentation can be deployed in the field setting, 

facilitating rapid screening analyses.

It is important to insure that the in-situ speciation of aqueous selenium is maintained following sample collection 

prior to analysis. Gómez Ariza et al. (2000) provide a 

recent review of studies conducted to assess appropriate 

methods for selenium species preservation in natural water 

samples. In general, this review suggests that acidification 

followed by sample storage at ≤4 °C is adequate to prevent 

changes in the oxidation state of selenite/selenate and/or 

selenium loss via coprecipitation with ferric (hydr)oxides 

due to air exposure over reasonable time periods. More 

stringent collection, handling, and storage requirements 

may be necessary for ground-water samples with significant 

concentrations of organoselenium species. Alternatively, 

it may be possible to carry out separation of the various 

selenium species within the field using solid phase extraction (SPE) immediately following sample collection. There 

are several published reviews of the use of solid phase 

extraction media that may be employed to selectively extract 

individual selenium species from water samples (Bueno et 

al., 2002; Camel, 2003; Wake et al., 2004). Subsequent 

analysis of the retained selenium species can then be 

achieved in the laboratory following elution of the species 

off the SPE material. This approach may be necessary 

for water samples containing selenium species that are 

not amenable to chemical preservation within the native 

ground-water sample.

Solid Phase Measurements

Determination of the chemical speciation of selenium within 

aquifer sediments may be required for determination of the 

mechanisms active in its sequestration from ground water. 

This analysis may include determination of the oxidation 

state of selenium within the whole solid matrix or individual 

components, as well as the determination of its presence 

as a discrete solid phase (e.g., elemental selenium) or 

its association with specific mineral components (e.g., 

iron oxides or sulfides). The following review provides a 

summary of approaches that have been used to assist in 

identification of the solid phase associations controlling 

selenium solubility within a contaminant plume.

There are several examples within the literature of the determination of the distribution of selenium oxidation states 

within a solid matrix following partial or complete dissolution 

of the solid (e.g., Ferri et al., 1998; Ochsenkuhn-Petropoulou and Tsopelas, 2002; de Leon et al., 2003). However, 

there has been insufficient assessment of the potential for 

selenium species transformation that may occur with use 

of aggressive extraction procedures designed to achieve 

complete sample dissolution (especially oxidation state) 

in order to recommend this analytical approach. Determination of selenium oxidation state for extracts that are 

non-selective for various mineralogical associations within 

the aquifer sediment may not provide sufficient detail to 

identify specific partitioning mechanisms. In these instances, the use of extraction solutions that are intended 

to target specific mineralogical associations may provide a 

necessary compliment to determination of the distribution 

of selenium oxidation states within the whole solid matrix. 

Again, there are numerous examples of the use of single 

or sequential extraction procedures to determine the mineralogical speciation of selenium in soils and sediments. 

The following provides a synopsis of the most commonly 

employed procedures.

Sequential extraction procedures (SEP) employed for 

characterization of aquifer sediments can be categorized 

into two types: 1) procedures designed to quantify the 

abundance of various mineralogical components (e.g., iron 

oxides, carbonates, organic matter), and 2) procedures 

designed to quantify the abundance of specific trace element phase associations. The first SEP types have a longer 

history of development and application and are considered 

as an appropriate supplement to other analytical methods 

(e.g., X-ray diffraction, thermal analysis) used to define soil/

sediment mineralogy. There are examples in the literature 

where this type of SEP has been used to infer selenium 

speciation in soil/sediments (e.g., Tokonaga et al., 1991). 

However, the chemical extractants employed in these procedures are not designed to specifically recover selenium 

from specific solid phase associations. In the absence of 

suitable standard reference materials (soils/sediments) 

and the general failure to include matrix spike analyses 

with reference compounds that represent various selenium 

solid phase species anticipated in soils/sediments (e.g., 

selenate/selenite adsorbed to reference soil minerals, elemental selenium, organoselenium compounds), the use of 

this approach is not likely to provide useful characterization 

information.

Sequential extraction procedures designed to target specific 

selenium solid phase associations (i.e., type two SEPs) have 

been developed and applied for characterization of soils and 

sediments. A recent study by Wright et al. (2003) provides 

a comprehensive review and analysis of the utility of these 

procedures. Ultimately, these authors demonstrated the 

non-selectivity of existing SEPs and provided analysis of an 

alternative SEP designed to target selenium in the following 

solid phase associations: 1) weakly adsorbed selenate, 2) 

strongly adsorbed selenite, 3) elemental selenium [Se(0)], 

and 4) selenium associated with the organic fraction. The 

authors provide a comprehensive analytical assessment 

of potential analytical artifacts and extraction phase selectivity through the use of model reference compounds 

incorporated into either a ‘synthesized’ reference soil or 

spiked into real soils/sediments for analytical performance 

analysis. Wright et al. (2003) point out that all procedures 

are not capable of uniquely identifying selenium in either 

an iron-selenide precipitate (e.g., FeSe(s)) or selenium incorporated into iron sulfides (e.g,, mackinawite or pyrite). 

Velinsky and Cutter (1990) suggest the use of extractants 
78

designed to target dissolution of iron sulfides as a means 

for assessing the amount of coprecipitated selenium, but 

this avenue requires additional testing to properly assess its 

reliability. Chu et al. (2006) also point to the potential use 

of carbon disulfide as a stand-alone extractant to target the 

most labile form of elemental selenium. Finally, it is also 

important to circumventing possible analytical artifacts at 

the stage of selenium detection in the extraction solution 

due to interferences from either 1) reagents employed in the 

extraction and sample pre-treatment for analysis (Wright et 

al., 2003) or 2) sample matrix components that were liberated from the soil or sediment (Bujdoš et al., 2000).

Recently, X-ray absorption spectroscopy/microscopy employing synchrotron radiation sources have been employed 

to provide direct assessment of the in-situ speciation of 

selenium in soils/sediments. The synchrotron X-ray fluorescence microprobe (SXRFM) provides the capability of 

mapping elemental distributions at a spatial resolution 

which, in some cases, approaches 1 μm. The application of 

SXRFM for mapping selenium distribution within both natural and laboratory-constructed soil aggregates illustrates 

the potential usefulness of this tool for site characterization 

(Tokunaga et al., 1994; Strawn et al., 2002). In addition, 

X-ray absorption spectroscopy has been employed to identify the in-situ chemical speciation of selenium (including 

selenium oxidation state and bonding environment) in both 

experimental and natural materials (Tokunaga et al., 1994; 

Strawn et al., 2002; Herbel et al., 2003; Templeton et al., 

2003). The advantage of these techniques is the ability 

to characterize the solid phase speciation of selenium at 

relatively low concentrations in soils/sediments with minimal sample handling that may perturb the characteristics 

of the as-collected sample. However, these methods are 

not routinely available, which limits their utility as a regular 

component of a site characterization effort. 

Selenium Isotope Fractionation

A review of the use of selenium isotope measurements for 

the determination of selenium reduction processes active 

within an aquifer or the determination of selenium sources 

within a plume is provided in Johnson and Bullen (2004). 

The six stable isotopes, 74Se, 76Se, 77Se, 78Se, 80Se, and 

82Se, are present in abundances of 0.889, 9.366, 7.635, 

23.772, 49.607, and 8.731%. Reduction of selenate and 

selenite, either through biotic or abiotic processes, may 

result in enrichment of lighter isotopes in the reduction 

products and a complementary enrichment of the heavier 

isotopes in the remaining unreduced selenium due to kinetic 

isotope fractionation processes (Johnson and Bullen, 2004). 

Since reduction of selenium results in breaking a bond with 

oxygen, the kinetic effect is a result of the smaller expense 

of energy by the microbe to reduce the lighter isotope. The 

current state of knowledge relative to the extent of selenium 

isotope fractionation as a result of abiotic-biotic oxidation/

reduction processes is summarized in Table 7.3 (Johnson 

and Bullen, 2004). This summary is in part derived from 

recent studies designed to assess the importance of biotic 

(Herbel et al., 2000; Ellis et al., 2003) and abiotic (Johnson 

and Bullen, 2003) reduction processes that are anticipated 

to be active in reducing subsurface environmentals. Currently available data indicate that reduction of the selenium 

oxyanions, selenate and selenite, is the main source of 

fractionation observed in ground-water systems.

Determination of selenium isotope fractionation can be carried out using either thermal ionization mass spectrometry 

(TIMS; Johnson et al., 1999) or multi-collector inductively 

coupled plasma mass spectrometry (MC-ICP-MS; Rouxel 

et al., 2002). Both methods are susceptible to measurement bias due to changes in selenium speciation and mass 

distribution that may occur during sample preparation due 

to the susceptibility of selenium to redox transformations 

and volatilization. Therefore, it is recommended that a 

‘double spike’ technique be employed to correct for these 

analytical inaccuracies (Johnson et al., 1999; Johnson 

and Bullen, 2004). With this method, a reference sample 

with known 82Se/74Se ratio is added to the sample and its 

recovery is monitored along with evaluation of the natural 

ratio of 80Se/76Se originally in the sample. In addition, the 

MC-ICP-MS method is also susceptible to bias due to the 

formation of polyatomic species with identical masses that 

form within the argon plasma (e.g., ArCl+ and ArC+; Rouxel 

et al., 2002). It should be noted that a representative interlaboratory standard is not commercially available, although 

the National Institute of Standards and Technology does 

market a reference meteorite standard (Canyon Diablo 

Table 7.3 Review of selenium isotope fractionation ranges for abiotic-biotic processes during reduction and oxidation (Johnson, 2004). NM = not measured

Reduction Abiotic Biotic Oxidation Abiotic Biotic

Se(VI) → Se(IV) 7-12 ‰ 3-5 ‰ Se(IV) → Se(VI) 0 ‰ NM

Se(IV) → Se(0) 6-13 ‰ 6-9 ‰ Se(0) → Se(IV) 0 ‰ NM

Se(0) → Se(-II) 0 ‰ NM Se(-II) → Se(0) Small NM
79

Troilite; SRM 3149). Typically, a provisional standard is 

generated via mixtures of synthetic solutions with known 

selenium isotope content, and this reference is used to 

monitor analytical performance. Isotopic measurements 

may be performed on the native ground water or chemical 

extracts used to determine the selenium isotope distribution 

in on whole or selective fractions of aquifer solids.

Long-term Stability and Capacity

The long-term stability of selenium immobilized onto aquifer 

solids will depend on the prevailing ground-water chemistry over time relative to the conditions that existed at the 

time of immobilization. If there are significant changes 

in ground-water chemistry following immobilization, then 

the potential exists for remobilization of selenium with the 

establishment of a new mobile plume. A general example 

of this type of situation would be if selenium was originally 

immobilized through partitioning to mineral sulfides within 

aquifer sediments under sulfate-reducing conditions. Selenium associated with sulfide minerals may be re-mobilized 

due to sulfide oxidation if the aquifer were to return to more 

oxidizing conditions. Alternatively, there is evidence that 

elemental selenium may be re-oxidized to selenite/selenate, 

resulting in re-mobilization of selenium in ground water 

(Tokunaga et al., 1996; Zawislanski and Zavarin, 1996; Losi 

and Frankenberger, 1998). However, as demonstrated by 

Dowdle and Oremland (1998) the oxidation of elemental 

selenium is a slow, microbially-controlled process with 

rates that are 3-4 orders of magnitude lower than microbial 

reduction of selenate/selenite.

The capacity to sustain continued removal of selenium from 

ground water will be dictated by the evolution of groundwater chemistry through time and the continued activity 

of relevant microbial processes. For situations in which 

adsorption or abiotic precipitation reactions control selenium attenuation, significant changes in pH or increases 

in the concentration of competing anions (e.g., sulfate) in 

ground water can cause reversal of adsorption reactions or 

dissolution of precipitates (e.g., metal selenates/selenites). 

Likewise, a change to more oxidizing conditions can cause 

destabilization of reduced mineral phases (e.g., sulfides) 

that may sequester selenium during the original attenuation reaction. For situations in which microbial reduction 

controls selenium attenuation, changes in the availability of 

required substrates/nutrients as well as potentially competing electron acceptors can reduce the capacity of the aquifer 

to sustain continued attenuation prior to consumption of the 

total mass of selenium within the plume. There is laboratory 

and field evidence that increases in ground-water nitrate 

can inhibit or slow down selenium reduction (Oremland et 

al., 1989; Benson, 1998). While these examples are not 

exhaustive of all possible situations, they do point to the 

importance of establishing a monitoring program that is 

designed to consider the potential impact of changes in 

site geochemistry relative to the conditions under which 

selenium is attenuated within the aquifer. Ultimately, the 

evolution of biogeochemical conditions in the subsurface 

will dictate the success of the MNA remedy.

Tiered Analysis

Determination of the viability of selenium remediation in 

ground water via monitored natural attenuation will depend 

upon proper assessment of contaminant loading to the 

aquifer and prevailing geochemistry and mineralogy within 

the contaminant plume and the down gradient zone prior 

to the point(s) of compliance. The goal of site assessment 

is to demonstrate the process(es) controlling selenium 

sequestration onto aquifer solids and the long-term stability of solid phase selenium as a function of existing and 

anticipated ground-water chemistry. The following tiered 

analysis structure for site characterization provides an approach to evaluate candidate sites and define the potential 

limitations of MNA as part of a remedy for ground-water 

cleanup.

Tier I Site characterization under Tier I will involve demonstration that the plume is static or shrinking, has not 

reached compliance boundaries, and does not impact 

existing water supplies. Once this is established through 

ground-water characterization, evidence is collected to 

demonstrate Se partitioning to aquifer solids within the 

plume. Rapid movement of contaminants along preferred 

flow paths in the unsaturated and saturated zones can be 

induced by hydrologic events such as heavy rains (e.g., 

McCarthy et al., 1998; Camobreco et al., 1996). It will be 

important to determine that such hydrogeologic features do 

not result in contaminants bypassing zones where natural 

attenuation is occurring. If natural attenuation processes 

are active throughout the plume, then an observed increase 

in solid phase concentrations within regions of the plume 

with higher aqueous concentrations is anticipated, e.g., 

near the source term. This field partitioning data may be 

supplemented by geochemical modeling that incorporates 

measured water chemistry (e.g., pH, Eh, and major ion 

chemistry) throughout the plume to assess the potential for 

solubility control by metal selenates/selenites or elemental 

selenium. Since identification of the chemical speciation of 

selenium in water samples and aquifer sediments is critical 

towards determining the attenuation mechanism(s), it is recommended that precautions be taken to preserve selenium 

speciation during collection, preservation, and processing 

of collected samples (See recommendations and application of these methods in Wilkin et al., 2002 and Ford et 

al., 2005). Identification of active sequestration to prevent 

selenium migration in ground-water provides justification 

for proceeding to Tier II characterization efforts.

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

of attenuation are determined. Estimates of a site attenuation rate(s) can be assessed via a well transect along 

the ground-water flow path. In addition, time-series data 

may be collected at one or more monitoring points within 

the plume. This information will allow assessment of the 

relative timescales for contaminant immobilization and fluid 

transport and determination of whether remediation objectives can be met within the required regulatory time frame. 

In addition, the mechanism(s) for attenuation need to be 

identified under this stage of site characterization. This 
80

effort may require determination of the chemical speciation 

of aqueous and solid phase Se, which may be approached 

according to the following scheme: 

1) Determination of selenium solution speciation 

via direct analytical measurements in combination with speciation calculations based on 

characterized ground-water chemistry;

2) Calculation of the saturation state of ground 

water relative to measured aqueous chemistry complimented by the possible isolation 

of discrete Se mineral phases via density 

separations (or other schemes) in regions of 

the aquifer with highest solid phase concentrations;

3) Determination of aquifer mineralogy to determine the relative abundance of components 

with documented capacity for Se sorption (e.g., 

Amonette, 2002);

4) Identification of selenium association(s) with 

the various solid phase components of aquifer 

solids through combination of chemical extractions with microscopic/spectroscopic confirmation of phase associations, and;

5) Demonstration of concurrence between the 

site conceptual model and mathematical 

model(s) that describe selenium removal 

mechanism(s).

It is recommended that identification of selenium chemical 

speciation in aqueous and solid matrices be conducted using samples collected in a manner that preserves the in-situ

mineralogy and speciation of selenium. The demonstration 

of concurrence between conceptual and mathematical models describing selenium transport will entail development 

of site-specific parameterization of the chemical processes 

controlling selenium solid phase partitioning. 

Tier III Once the attenuation mechanism(s) has been 

identified for the site, the subsequent characterization effort 

under Tier III will involve determination of the stability of 

immobilized Se and the capacity of the aquifer to sustain 

continued uptake. It is recommended that the stability of 

immobilized Se be tested based on the anticipated evolution 

of ground-water chemistry concurrent with plume shrinkage. 

For example, changes in ground-water redox potential due 

to oxygen intrusion or the occurrence of aerobic microbial 

processes can exert a significant influence on Se attenuated 

in its elemental form [Se(0)]. Therefore, it is recommended 

that sediment leach tests be conducted to characterize the 

magnitude of Se mobilization as a function of ground-water 

redox chemistry representative of existing and anticipated 

site conditions. This may involve the use of microcosm 

tests that stimulate in-situ microbial populations toward the 

development of redox conditions considered deleterious for 

continued Se immobilization. It is recommended that the 

capacity for Se uptake onto aquifer solids be determined 

relative to the specific mechanism(s) identified in Tier II. 

For example, if site characterization under Tier II indicated 

that co-precipitation of Se with iron sulfide due to microbial 

degradation of organic compounds coupled with sulfate 

reduction occurs within the aquifer, it is recommended 

that the mass distribution of organic carbon, sulfate and 

ferrous iron to support this reaction within the aquifer 

be determined. This site-specific capacity can then be 

compared to Se mass loading within the plume in order to 

assess the longevity of the natural attenuation process. If 

site-specific tests demonstrate the stability of immobilized 

Se and sufficient capacity within the aquifer to sustain Se 

attenuation, then the site characterization effort can progress to Tier IV. For cases where contaminant stability is 

sufficient but aquifer capacity is insufficient for capture of 

the entire plume, then a determination of the benefits of 

contaminant source reduction may be necessary. 

Tier IV Finally, under Tier IV a monitoring plan is established along with contingency plans in the event of MNA 

failure. It is recommended that the monitoring plan be 

designed to establish both continued plume stability and to 

identify changes in ground-water chemistry that may lead 

to re-mobilization of attenuated Se. The specific chemical 

parameters to be monitored will include those identified 

under Tier III that may halt Se partitioning to aquifer sediments and/or result in solubilization of either discrete Se 

precipitates or aquifer minerals that sequester Se from 

ground water. For example, solution phase parameters 

that could alter either Se precipitation or adsorption include 

increases in the concentration of competing anions, such 

as sulfate, in combination with changes in ground-water 

pH. In contrast, the concentration of dissolved iron or sulfate may indicate the dissolution of an important sorptive 

phase within the aquifer (e.g., reductive dissolution of iron 

oxyhydroxides or oxidative dissolution of sulfides). Changes 

in these parameters may occur prior to observed changes 

in solution Se and, thus, serve as monitoring triggers for 

potential MNA failure. In this instance, a contingency plan 

can be implemented that incorporates strategies to arrest 

possible plume expansion beyond compliance boundaries. 

Possible strategies to prevent plume expansion include 

pump and treat operations, installation of reactive barriers 

to enhance uptake capacity perpendicular to the direction 

of plume advance, or enhancement of natural attenuation 

processes within the aquifer through the injection of soluble 

reactive components.

References 

Abu-Erreish, G.M., E.I. Whitehead, and O.E. Olson. Evolution of volatile selenium from soils. Soil Science 106: 

415-420 (1968).

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87

Nitrate 

Chunming Su, Robert G. Ford

Occurrence and Distribution

Nitrate (NO3

-

) is a chemical form of nitrogen that is essential 

for plant growth, but potentially toxic to human and animal 

life at moderate concentrations. Nitrate is considered to 

be the most ubiquitous chemical contaminant in aquifers 

throughout the world with continuously observed increases 

in contamination (Spalding and Exner, 1993). Results from 

the USGS National Water Quality Assessment Program 

showed that in the United States nitrate is detected in 71% 

of ground water samples, more than 13 times as often as 

ammonium (NH4

+), nitrite (NO2-

), organic nitrogen, and 

orthophosphate, based on a common detection threshold 

of 0.2 mg of N L-1 (Nolan and Stoner, 2000). According 

to these authors, shallow ground water (< 5 m deep) 

beneath agricultural land has the highest median nitrate 

concentration (3.4 mg of N L-1), followed by shallow ground 

water beneath urban land (1.6 mg of N L-1) and deeper 

ground water in major aquifers (0.48 mg of N L-1). Nitrate 

exceeds the maximum contaminant level (MCL), 10 mg of 

N L-1, in more than 15% of ground-water samples from 4 

of 33 major aquifers commonly used as a source of drinking water (Nolan and Stoner, 2000). Recent analyses of 

ground-water samples from about 1500 domestic drinking 

water and public supply wells show that of more than 140 

contaminants measured, nitrate most frequently exceeded 

drinking water standards or human health criteria (Squillace et al., 2002). 

A National Research Council report has indicated that there 

are approximately 300-400 thousand nitrate-contaminated 

sites in the United States (National Research Council, 

1994). Sources of nitrate and nitrite in ground water include 

atmospheric deposition from fossil fuel burning, runoff from 

fertilizer use; leaching from animal wastes from confined 

animal feedlot operations and dairies, septic tanks, and 

sewage; solid waste disposal (landfills and waste tips); 

and erosion of natural deposits (Puckett, 1995; Nolan and 

Stoner, 2000). Organic nitrogen and ammonia are potential 

nitrate sources because they tend to be converted to nitrate 

in natural waters. 

While nitrate leaching from disturbed forests is a threat to 

ground water, this source is small compared to agricultural 

and industrial sources (Keeney, 1986). A survey of eastern 

U.S. watersheds showed that the total nitrogen levels in 

streams draining agricultural watersheds were fivefold 

greater than forested watersheds, while the percentage 

of total nitrogen was much greater from agricultural than 

from forested watersheds (Omernik, 1976). Since most 

native and extensively managed grasslands are nitrogen 

deficient, very little nitrate will be available for leaching; also 

many of these grasslands are in the semiarid West where 

leaching is limited. On the other hand, intensively managed 

forage and grazed grasslands (with major nitrogen inputs 

from fertilizer and symbiotic nitrogen fixation) may be the 

source of considerable ground water nitrate. Grasslands, 

like croplands, have annual above-ground biomass cycles 

that leave nitrate in the soil profile susceptible to leaching 

at times of the year when plant uptake is minimal, usually 

spring and autumn (Keeney, 1986). Agricultural cropland 

provides a large nonpoint source relative to other nitrate 

sources. Nitrogen applications in amounts exceeding 

the optimum rates are often used to provide maximum 

economic yields. Excess nitrate in the root zone is leached 

at times when the soil is vulnerable to substantial rainfall 

or excessive irrigation. Irrigation combined with high use of 

nitrogen fertilizers is the primary source of nitrate in ground 

water (Keeney, 1986). 

Most of the areas high in ground-water nitrate are west of 

the Missouri River where irrigation is a necessity. Aquifers 

in highly agricultural areas in the southeastern USA reportedly are not contaminated. Vegetative uptake and denitrification in this warm, wet, carbon-rich environment are 

responsible for the natural remediation of nitrate in shallow 

aquifers (Spalding and Exner, 1993). In the Middle Atlantic 

States and the Delmarva Peninsula, localized contamination occurs beneath cropped, well-drained soils that receive 

excessive applications of manure and commercial fertilizer. 

Extensive tile drainage has for the most part prevented a 

nitrate problem in the ground water of the Corn Belt states 

in that the nitrate-contaminated recharge is diverted by tile 

drains and subsequently is discharged to surface water 

(Spalding and Exner, 1993). 

Point-source nitrate contamination includes ammonium 

nitrate as explosives residues in mining operations (e.g., 

Johnson et al., 2000). Nitrate in ground water can also 

arise from deposits lain down during geological times. Examples are nitrate found in Pleistocene age loess of semiarid southwestern and western central Nebraska (Boyce 

et al., 1976) and high levels of geological nitrogen found 

in the alluvium beneath the San Joaquin Valley, California 

(Strathouse et al., 1980; Holloway et al., 1998). A large 

reservoir of bioavailable nitrogen (up to 104 kilograms 

of nitrogen per hectare, as nitrate) has been previously 
88

overlooked in studies of global nitrogen distribution (Walvoord et al., 2003). The reservoir has been accumulating 

in subsoil zones of arid regions throughout the Holocene. 

Consideration of the subsoil reservoir raises estimates of 

vadose-zone nitrogen inventories by 14 to 71% for warm 

deserts and arid shrublands worldwide and by 3 to 16% 

globally. Subsoil nitrate accumulation indicates long-term 

leaching from desert soils, impelling further evaluation 

of nutrient dynamics in xeric ecosystems. Evidence that 

subsoil accumulations are readily mobilized raises concern 

about ground-water contamination as a result of land-use 

or climate change.

Ground water, due to its direct connection to surface waters, can be a significant source of nitrogen to lakes, impoundments, and estuaries. Increased nitrogen input from 

ground-water sources has caused overgrowth of aquatic 

plants and subsequently depletion of dissolved oxygen 

as plants decay (eutrophication). Although overlooked in 

many nutrient management plans, ground-water discharges 

of nitrate to streams may contribute the majority of the 

total amount of stream nitrogen (Williamson et al., 1998). 

Ground-water contamination at landfill sites by ammonium 

can be significant with mean concentrations of about 900 

mg NH4

+-N L-1 having been reported for landfill leachates 

in the United Kingdom (Burton and Watson-Craik, 1998). 

Ammonium production is a result of amino acid fermentation 

during decomposition of organic matter in domestic waste 

disposed under anaerobic conditions. Anaerobic conditions 

in the waste may persist for many years particularly if the 

landfill is capped with a low permeability cover to prevent 

water infiltration and gas exchange. Ammonium attenuation in subsoils and ground water is predominantly due to 

cation exchange and/or nitrification (biological oxidation) 

processes (Buss et al., 2004). Currently, more research is 

being conducted to improve the understanding of nitrogen 

cycling at the watershed level and to determine where 

nitrogen attenuation is occurring and to what extent. This 

information is needed to assist the U.S. Environmental 

Protection Agency and States with the general development 

of total maximum daily load (TMDL) for nitrogen, as well 

as to help States adopt more appropriate nutrient criteria 

on which TMDLs are based.

Plume Characteristics

Nitrate concentration in ground water is variable and depends on interactions among several factors, including 

nitrogen loading, soil type, aquifer permeability, recharge 

rate, climate, and aquifer oxidation state and pH. Factors 

that generally increase nitrate concentrations in ground 

water include well-drained soils, fractured bedrock, and 

irrigation. Factors that mitigate nitrate contamination of 

ground water include poorly drained soils, greater depth 

to ground water, artificial drainage systems, intervening 

layers of unfractured bedrock, a low rate of ground water 

recharge, and anaerobic conditions in aquifers (Nolan and 

Stoner, 2000). Vulnerability of ground water to contamination by nitrate does not depend on any single factor, but on 

the simultaneous influence of factors representing nitrogen 

loading sources and aquifer susceptibility (Nolan, 2001). 

Tools have been developed to assess aquifer susceptibility to nitrate contamination based on the factors indicated 

above (Ceplecha et al., 2004). Research conducted within 

the North Carolina Coastal Plain illustrates the dependence of nitrate transport on the variability of subsurface 

hydrogeologic characteristics. The North Carolina Coastal 

Plain consists of varying permeability surface sediments 

underlain by relatively impermeable, highly reduced sediments (Gilliam et al., 1974). Nitrate accumulates in the 

shallow ground water in response to agricultural activities, 

but because of the reduced zones (up to 15 m thick) it 

does not reach deep aquifers (Gilliam et al., 1979). It was 

found that if the water table was artificially raised using 

flashboard riser-type water-level control structures on tile 

mains or outlet ditches, denitrification in the subsoils could 

be enhanced to reduce nitrate concentrations in shallow 

ground water (Skaggs and Gilliam, 1981).

Remedial Technologies

A recent review of available technologies for remediation 

of nitrate contamination in ground water has been 

published by the Interstate Technology Regulatory Council 

(ITRC, 2000), which is available at http://www.itrcweb.

org/Documents/EISBD-1.pdf. Available technologies 

include: 1) ex-situ, pump-and-treat approaches that 

extract nitrate from recovered ground water or degrade 

nitrate to innocuous compounds through above-ground 

biotic or abiotic chemical reduction processes, and 2) insitu treatment approaches that degrade nitrate within the 

plume via phytoremediation, stimulation of native microbial 

communities for denitrification, or the emplacement of 

media that achieve biotic and abiotic chemical reduction 

of nitrate. Examples of field demonstrations of biotic 

remediation technologies for removal of nitrate from 

ground water are listed in Table 8.1. In addition, research 

indicates that abiotic processes may be used to remediate 

nitrate via ex-situ or in-situ approaches. Technologies 

that extract nitrate from water via concentration onto solid 

media (e.g., Tezuka et al., 2004) or into a waste brine 

have the disadvantage of requiring further treatment or 

disposal of the process waste stream (e.g., Dorsheimer 

et al., 1997). Examples of chemical reduction of nitrate 

via systems with reduced Fe media (Chew and Zhang, 

1998, 1999; Alowitz and Scherer, 2002; Chen et al., 2004; 

Su and Puls, 2004; Chen et al., 2005; Mishra and Farrell, 

2005; Yang and Lee, 2005) indicate that the predominant 

reaction leads to the production of ammonia-nitrogen, 

although the potential exists to control reaction conditions 

to favor complete nitrate reduction to dinitrogen gas. There 

is evidence that zero-valent iron may also stimulate the 

growth of subsurface denitrifying bacteria (Gu et al., 2002), 

suggesting the use of this material in permeable reactive 

barriers for mixed contaminant plumes containing nitrate 

and other inorganic/organic constituents more amenable 

to chemical reduction.

The success of these treatment approaches will be influenced by the level of knowledge of the ground-water flow 

regime as well as the geochemical characteristics within 

the plume. For technologies based on nitrate extraction, the 
89

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

Technology Application Setting Approach Reference

Phytoremediation/ Riparian Buffers/ Constructed Wetlands

Prevent deep percolation of nitrate at an abandoned feedlot high in soil profile nitrate

Planted corn (Zea mays L.) and deep-rooted 

alfalfa (Medicago sativa L.)

Schuman and Elliott (1978)

Cover cropping in the Northern U.S. Corn Belt to decrease nitrate loss from subsurface drainage discharge Planted autumn-seeded rye (Secale cereale L.) following corn harvest Strock et al. (2004)

Desert soil at a former uranium ore-processing plant, near Monument Valley, Arizona

Planted fourwing saltbush (Atriplex canescens); biological denitrification McKeon et al. (2005)

Nutrient interception by a riparian forest adjacent to cropland on Delmarva peninsula

Monitoring network consisting of well transect from crop field through hill slope, adjacent flood plain, and into receiving stream; measured nitrate removal

Jordan et al. (1993)

Wastewater effluent flowed through shallow ground water of a forested headwater riparian zone within the Appalachian Valley and Ridge physiographic province

Monitored denitrification enzyme activity attributable to riparian zone location, soil horizon, and nitrate amendments

Flite et al. (2001)

In-situ treatment; Permeable Reactive Barriers

Permeable reactive barriers to intercept effluent from septic system infiltration beds, horizontally migrating septic system plume, farm field drainage water, or a shallow sand-and-gravel aquifer

Porous media barrier composed of waste cellulose solids (wood mulch or particles, sawdust, leaf compost); horizontal infiltration beds in vadose zone and vertical wall or buried containerized systems in saturated zone

Robertson and Cherry (1995); Robertson et al. (2000); Robertson et al. (2005)

Permeable reactive subsurface barriers at two field sites in an agricultural area of southwestern Ontario

Upflow reactors with wood particle carbon source with very high hydraulic conductivities (0.1–1 cm s–1)

van Driel et al. (2006)

In-situ treatment; Permeable Reactive Zones via Injection

Stimulation of in-situ biological denitrification via subsurface injection of a carbon source such as ethanol or acetic acid

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

Hiscock et al. (1991); Hamon and Fustec (1991)

Plumes originating from former waste disposal ponds at the Y-12 plant Complex within the Oak Ridge National Laboratory; acidic sediments, low ground-water pH (pH 3.8 – 4.4), high nitrate (9-154 mM)

Conducted push-pull tests in wells to stimulate and monitor denitrification; injected solutions consisted of neutralized site ground water made anoxic by sparging with 80%N2-20%CO2, and 

amended with either ethanol or glucose as an electron donor 

North et al. (2004)

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

Sodium formate solution injected as a potential electron donor for denitrification; nitrate reduction occurred with buildup of nitrite due to incomplete reduction Smith et al. (2001)
90

presence of common anions in ground water may reduce 

treatment performance for engineered systems based on 

ion exchange, reverse osmosis, or electrodyolisis. For 

microbial nitrate reduction (ex-situ or in-situ), geochemical 

parameters such as low pH and high total dissolved solids, 

as well as the presence of aerobic conditions or competing electron acceptors such as perchlorate or sulfate may 

negatively impact the rates and extent of nitrate reduction 

that can be achieved. It is important to consider these sitespecific factors when selecting treatment technologies.

Regulatory Aspects

Nitrate was ranked number 216, nitrite number 212, and 

ammonia number 160 out of a total of 275 contaminants 

on the CERCLA Priority List of Hazardous Substances in 

2001 (Centers for Disease Control, 1996). The USEPA has 

set the MCL for nitrate at 10 mg of N L-1 (USEPA, 2006; 

http://www.epa.gov/waterscience/criteria/drinking/dwstandards.pdf). The USEPA has determined that infants below 

the age of six months who drink water containing nitrate 

or nitrite in excess of the MCL could become seriously ill 

and, if untreated, may die. Symptoms of acute exposure 

include shortness of breath and blueness of the skin, or 

blue-baby syndrome (methemoglobulinemia). This is due to 

the conversion of nitrate to nitrite by the body, which interferes with the oxygen-carrying capacity of the child’s blood. 

Chronic exposure to high levels of nitrate/nitrite results in 

diuresis, increased starchy deposits, and hemorrhaging of 

the spleen (http://www.epa.gov/safewater/contaminants/

dw_contamfs/nitrates.html). 

Geochemistry and Attenuation Processes

Aqueous Speciation

Nitrate (with an oxidation number of +5 for nitrogen) is the 

conjugate base of HNO3, a strong mineral acid. Common 

metal nitrate salts are highly soluble in water. Nitrate forms 

weak coordination complexes with metals and displays 

minimal sorption to inorganic or organic surfaces. Being an 

anion, nitrate (NO3

-

) in solution is attracted electrostatically 

to positively charged surfaces and is repulsed by negative 

surfaces. Nitrate is not appreciably adsorbed onto aquifer 

sediments. 

Redox Chemistry

Denitrification refers to a microbial respiratory process 

where nitrate is used as a terminal electron acceptor and 

is reduced to dinitrogen gas by the following generalized 

half reaction:

 2NO3

-

 + 12 H+ + 10e → N2 + 6H2O (Eq. 8.1)

It is an important biological process for the return of fixed 

nitrogen to the atmosphere. In the above respiratory electron transport process electron donors are usually organic 

matter or reduced sulfur compounds. Nitrate can also be 

reduced to ammonium, but this requires highly reducing 

conditions and organic carbon. The denitrifying bacteria 

use nitrate, nitrite, or nitrous oxide as terminal electron acceptors for respiratory growth in the presence of gaseous 

oxygen and under micro-aerobic and anaerobic conditions. 

Denitrification involves four consecutive steps in which nitrate is reduced to dinitrogen gas by the metalloenzymes, 

nitrate reductase, nitrite reductase, nitric oxide reductase, 

and nitrous oxide reductase:

NO3

- → NO2-

 → NO → N2O → N2 (Eq. 8.2)

Many denitrifying organisms do not contain or express all 

of the four reductases needed for complete denitrification 

so that the interaction of different organisms is needed to 

reduce nitrate completely into dinitrogen gas.

 Anaerobic conditions in the immediate cell environment 

are essential for denitrification to occur (Table 8.2). This 

translates to a measured Eh potential of 300 to 350 mV 

(Reddy and Patrick, 1984). Denitrification is not limited to 

water-logged soils, it occurs in moist, fine-textured soils as 

well as in localized micro-anoxic zones within the overall 

aerobic soil matrix. The optimum temperature range for 

denitrification is from 20 to 35 °C and the optimum pH is 

neutral to slightly alkaline, but denitrification will occur over 

at least pH 5 to 9 (Focht and Verstraete, 1977). Denitrification can be described by a zero-order kinetic reaction at 

high concentrations of nitrate, and by a first-order reaction 

at low concentrations (Focht and Verstraete, 1977).

There are two physiologically distinct classes of denitrifying bacteria: heterotrophic denitrifiers, which use organic 

Table 8.2 Natural attenuation and mobilization pathways for nitrate.

Attenuation 

Processes Mobilization Processes Characterization Approach

Biotic (microbial) or 

abiotic transformation of nitrate to other 

nitrogen species.

Cessation of microbial processes via 

changes in nutrient supply within the 

ground-water flow path; reduction 

in the mass of reduced Fe-bearing 

minerals within the aquifer due to 

changes in redox conditions or supply 

of soluble iron.

Evaluation of N speciation in ground water. Determination of water chemistry, microbial populations, 

and/or sediment mineralogy linked to nitrate transformation along with spatial and temporal variability 

of required chemical/microbial components relative 

to nitrate transport pathway(s).
91

compounds as electron donors and carbon sources (Equation 8.3 and 8.4); and chemolithoautotrophic (or autotrophic) 

denitrifiers, which use inorganic compounds (e.g., minerals 

containing reduced sulfur and iron) as electron donors and 

CO2 as a carbon source (Equations 8.5 and 8.6):

Heterotrophic denitrification:

4 NO3

- + 5CH2O (organic matter) →

2 N2 + 5HCO3

-

 + H+ + 2H2O (Eq. 8.3)

8 NO3

- + 5CH3COO-

(acetate) + 3H+ →

4 N2 + 10HCO3

-

 + 4H2O (Eq. 8.4)

Autotrophic denitrification:

14 NO3

-

 + 5FeS2 (pyrite) + 4H+ →

7 N2 + 5Fe2+ + 10 SO4

2- + 2H2O (Eq. 8.5)

NO3

-

 + 5Fe2+ + 7H2O →

5FeOOH (goethite) + 0.5 N2 + 9H+ (Eq. 8.6)

Because the rate and extent of denitrification are dependent 

on available carbon supply (Reddy and Patrick, 1984), it 

is often assumed that denitrification is not important as a 

nitrate-removal mechanism below the root zone. Denitrification is usually not significant in the subsurface of low 

organic matter and of sandy soils. 

Nitrate was found in the ground water beneath poorly 

drained soils of the North Carolina Coastal Plain and denitrification was proposed as a nitrogen loss mechanism in 

the saturated subsoil (Gilliam et al., 1974). Based on nitrate 

concentration data, nitrogen mass balances, soluble organic 

carbon, and redox potential measurements, Gambrell et al. 

(1975) concluded that denitrification could readily occur in 

the subsurface (1 to 2.5 m) of poorly drained subsoils of 

the North Carolina Coastal Plain. They attributed this to 

the high amounts of dissolved organic carbon (the subsoil 

was overlain by an organic soil), low dissolved oxygen, 

and low Eh (350 mV at 1 m). The nitrate concentration and 

nitrate-to-chloride concentration ratios declined sharply 

with depth. The North Carolina case is relevant to many 

other riparian and wetland areas that are believed to be 

significant in denitrification losses at the watershed level. 

Under certain flow and redox conditions, riparian zones 

along waterways have demonstrated a natural proficiency 

to intercept and denitrify nitrate in shallow ground water 

(Gilliam, 1991; Schipper et al., 1991).

The rate and mechanism of nitrate removal along and between ground water flow paths were investigated using a 

series of well nests screened in an unconfined sand and 

gravel aquifer in southwestern British Columbia, Canada 

(Tesoriero et al., 2000). Intensive agricultural activity in this 

area has resulted in nitrate concentrations in ground water 

often exceeding drinking water standards. Both the extent 

and rate of denitrification varied depending on the ground 

water flow path. While little or no denitrification occurred 

in much of the upland portions of the aquifer, a gradual 

redox gradient was observed as aerobic upland ground 

water moved deeper into the aquifer. In contrast, a sharp 

shallow redox gradient was observed adjacent to a thirdorder stream as aerobic ground water entered reduced 

sediments. An essentially complete loss of nitrate concurrent with increases in excess dissolved N2 gas provided 

evidence that denitrification occurred as ground water 

entered this zone. Low denitrification rates were observed 

along the deep flow path (< 0.04 μmol cm-3 aquifer yr-1). 

The estimated denitrification rate for the redoxcline in this 

aquifer ranged from 0.3 to 0.8 μmol cm-3 aquifer yr-1, and 

an in-situ experiment conducted adjacent to the stream 

suggested potential denitrification rates may be as high as 

42 μmol cm-3 aquifer yr-1. Electron and mass balance calculations suggested that iron sulfide (e.g., pyrite) deposits 

and to a lesser degree organic matter were electron donors 

for denitrification. 

Shallow anaerobic ground water not immediately adjacent 

to streams exhibits variable but generally low rates of denitrification. Natural gradient tracer tests yielded potential 

denitrification rates ranging from 0.44 to 1.1 μmol cm-3

aquifer yr-1 for a shallow sand and gravel aquifer on Cape 

Cod, Massachusetts (Smith et al., 1996). Greater denitrification rates (24 μmol cm-3 aquifer yr-1) were estimated 

from an in-situ experiment conducted 3 m below the land 

surface in a shallow aquifer in Ontario, Canada (Trudell et 

al., 1986). 

Pyrite-bearing aquifers represent important hydrological 

compartments due to their capacity to eliminate nitrate. In 

the absence of molecular oxygen, the nitrate is reduced 

coupled with the microbial oxidation of sulfur in the pyrite 

by the bacterium Thiobacillus denitrificans (Pauwels et al., 

1998, Grimaldi et al., 2004). A field tracer test using Brcoupled with nitrate injection revealed that denitrification 

rates in a pyrite-bearing schist aquifer varied depending 

on the medium permeability, with slower the denitrification 

rate observed in regions of faster ground-water flow velocity 

(Pauwels et al., 1998). Denitrification was shown to be accompanied by the precipitation of sulfate and iron-bearing 

minerals, probably jarosite (KFe3(SO4)2(OH)6). 

Under anaerobic conditions, the presence of both nitrate 

and ammonium may facilitate establishment of anaerobic 

ammonium-oxidizing (anammox) bacteria that directly 

oxidize ammonium to dinitrogen gas with nitrate (Mulder 

et al., 1995) and nitrite (Dalsgaard et al., 2003) as the 

electron acceptor:

5NH4

+ + 3NO3-

 → 4N2 + 9H2O + 2H+ (Eq. 8.7)

NH4

+ + NO2-

 → N2 + 2H2O (Eq. 8.8)

It remains to be seen if ammonium accumulation is an issue 

of concern where nitrate is a major remediation target.

Chemical reduction of nitrate has been shown to occur in 

a mixture of dissolved Fe2+ and Fe(II)-containing silicates 

(arfvedsonite and augite), but only after an oxyhydroxide 

precipitate had formed on the silicate surface (Postma, 

1990). Similarly, nitrite has been shown to be reduced 

to N2O by adsorbed Fe(II) on the surface of lepidocrocite 

(γ-FeOOH) at pH > 7 (Sørensen and Thorling, 1991). The 

Fe(II) associated with a reactive complex formed during 

Fe2+ binding to lepidocrocite, and not the ionic Fe2+, seemed 

responsible for the reduction of NO2

-

 similar to the follow-
92

ing reaction, where the first term in the reaction signifies 

ferrous iron adsorbed to lepidocrocite.

6Fe2+

(ads) + 2 NO2

- + 5H2O →

2Fe3O4(s) + N2O + 10H+ (Eq. 8.9)

In this reaction, it is presumed that the ferrous iron-lepidocrocite surface complex is converted to surface precipitate 

similar in nature to the mineral magnetite. The catalytic 

effect of Fe(III) oxyhydroxide may stimulate Fe(II)-dependent formation of N2O from NO2

-

 (chemodenitrification) in 

sediments and subsoils.

In sulfate green rust [Fe(II)4Fe(III)2(OH)12SO4• nH2O] suspensions, nitrate is stoichiometrically reduced to NH4

+, 

and magnetite is the sole Fe-containing product (Hansen 

et al., 1996).

4Fe(II)4Fe(III)2(OH)12SO4 + NO3

-

 + 6OH- →

NH4

+ + 8Fe3O4 + 4SO42- + 24H2O (Eq. 8.10)

Recent thermodynamic and spectroscopic studies give 

direct evidence for the existence of green rusts in soils 

(Hansen et al., 1994; Trolard et al., 1997). Since microbial 

processes may be expected only to dominate in environments rich in organic carbon, the green rust-facilitated NH4+ 

formation may be an important pathway for nitrate removal 

and nitrogen conservation in certain anoxic subsoils and 

sediments poor in organic matter.

Site Characterization

Groundwater geochemistry provides information about the 

source, transformation, and attenuation of nitrate in ground 

water. In addition, specific soil properties are important 

parameters determining the potential for natural attenuation 

of nitrate. For example, nitrate concentrations at the 0.8 to 

8 m depth at 15 sites within a beef feedlot were found to 

decrease as clay content increased (Lund et al., 1974). 

Other characteristics, such as a clay layer that severely 

restricted water movement, also decreased soil nitrate 

concentrations. It was probable that considerable soluble 

carbon was leached from the feedlot floor to provide energy 

for denitrification. The presence of ample organic carbon in 

the soil leads to active denitrification. Studies on the fate of 

nitrate below organic waste disposal sites (food processing 

waste lagoons, manure lagoons, feed yard, sewage sludge, 

and effluent septic tank drainfields) have shown significant 

denitrification in the vadose zone (Keeney, 1981). 

Measurement of Inorganic Nitrogen Species

It is recommended that site characterization include 

measurement of the various inorganic nitrogen species 

– largely nitrate, nitrite, and ammonia in ground water. At 

circum-neutral pH typical of most ground water, ammonia 

is present predominantly as the ammonium ion (NH4

+). 

Nitrogen is a redox sensitive element that is involved 

in numerous chemical and biological processes. It is 

anticipated that the fate and transformation of nitrate, 

nitrite, and ammonia in the subsurface would be reflected 

by the changes of their concentrations in the aqueous 

phase of the subsurface environment with time. It is 

therefore important to monitor their concentrations as an 

integral part of the overall evaluation of natural attenuation 

at a site. A summary of available laboratory methods for 

determination of the various inorganic species of nitrogen 

in ground water is provided in Table A.1 in USEPA (2002), 

which is available at http://www.epa.gov/ada/download/

reports/epa_600_r02_002.pdf. 

Although field colorimetric methods are available to quantify nitrate, nitrite, and ammonia, field determination using 

these methods is not generally needed as long as laboratory measurements can be made within specified holding 

times. Samples collected for nitrogen speciation are to 

be preserved by keeping cold (4°C) and measured within 

specified holding times, generally < 48 h. Depending on 

the method of analysis, longer holding times are possible 

by acidifying samples with sulfuric acid and keeping them 

cold. Prior to analysis, acid-preserved samples are to be 

brought back to room temperature and neutralized by adding base.

Use of Nitrogen Isotope Ratios for 

Identifying Sources and Transformation 

Processes of Nitrate

There are two stable isotopes of nitrogen: 14N (with a natural 

abundance of 99.6337 atom percentage) and 15N (0.3663) 

(Junk and Svec, 1958). Enrichment of 15N is expressed on 

the per mille (‰) basis, and is calculated as:

δ15N‰ = (atom%15Nsample – atom%15Nstandard) × 

1000 / (atom%15Nstandard) (Eq. 8.11)

The standard is usually atmospheric nitrogen. Manufactured fertilizers and nitrate in precipitation tend to have 15N 

concentrations close to natural abundance (δ15N‰ = 0), 

and soil organic matter nitrogen and hog manure organic 

nitrogen show greater δ15N‰ values (Kellman, 2005). In 

bacterially mediated processes the reaction of the light 

isotopic species is kinetically favored. As a result, the 15N/14N 

ratio of the instantaneously-formed product is lower than 

the 15N/14N ratio of the remaining, unreacted substrate. 

Ammonia volatilization can also lead to 15N isotope enrichment in the remaining substrate. This “kinetic isotope 

fractionation” is defined in terms of the isotope enrichment 

factor, εproduct-substrate (εp-s), 

εp-s ≈ δ15Nproduct - δ15Nsubstrate (Eq. 8.12)

For bacterial processes, εp-s is negative or zero (Heaton 

et al., 2005). If a process can be approximated as a first 

order reaction, changes in the δ15N value of the product or 

substrate of the reaction can be calculated using “Rayleigh” 

equations (Heaton et al., 2005).

For the substrate,

δs, t = δs, 0 + εp-s.ln(f), (Eq. 8.13)

and for the total accumulated product,

δp, t = δs, 0 - εp-s.ln(f)[f/(1-f)] (Eq. 8.14)

Where δs, t, δp, t are the δ15N values of the remaining substrate and accumulated product at time t; δs, 0 is the initial 
93

δ15N value of the substrate; f is the fraction of substrate 

remaining; and εp-s is the isotope enrichment factor for the 

process.

Since the early 1990s, a few published studies have focused 

on the transport and fate of nitrate in ground water using 

dual isotope analysis (e.g., δ15N and δ18O in dissolved NO3

-

) 

in combination with other lines of evidence to document 

in-situ microbial denitrification (Aravena and Robertson, 

1998; Beller et al., 2004). Analytical procedures have been 

developed for determination of δ15N via catalytic reduction of 

nitrate to N2 after anion exchange separation and precipitation as AgNO3, and for determination of δ18O via combustion of nitrate to CO2 using excess graphite (Aravena and 

Robertson, 1998). For determination of isotopic signatures 

of both nitrate and nitrite, a quantitative method has been 

developed based on bacterial denitrification (Sigman et 

al., 2001; Casciotti et al., 2002). Widory et al. (2005) used 

a coupled isotopic approach (δ15N and δ11B), in addition to 

conventional hydrogeological analyses, to trace the origin of 

NO3

-

 in ground water. The studied watersheds include both 

fractured bedrock and alluvial (subsurface and deep) hydrogeological contexts. The joint use of nitrogen and boron 

isotope systematics in each context deciphers the origin of 

NO3

-

 in the groundwater and allows a semi-quantification of 

the contributions of the respective pollution sources (mineral 

fertilizers, wastewater, and animal manure).

To investigate the fate of nitrate in a petroleum-contaminated aquifer, 15N isotope and acetylene-inhibition methods 

in combination with single-well push-pull tests were used 

to quantify processes contributing to nitrate consumption 

(Schurmann et al., 2003). The processes quantified included 

denitrification, assimilatory nitrate reduction, dissimilatory 

nitrate reduction to ammonium, and abiotic nitrate reduction. Multiple lines of evidence from chemical composition, 

stable isotope (13C/12C, 15N/14N, and 34S/32S), and dissolved 

gas (N2, Ar, O2, and CH4) composition of ground water at a 

landfill site were used to account for all NH4

+ loss by combined nitrification and denitrification processes in a system 

where there were abrupt temporal and spatial changes in 

redox conditions (Heaton et al., 2005). 

There are pitfalls in nitrogen source identification using the nitrogen isotope approach. The heterogeneity of 

complex watersheds makes source identification difficult; 

the complexity of the nitrogen cycle and the associated 

isotope fractionations can make the results ambiguous or 

unreliable; source identification becomes very difficult when 

mixing of point and non-point sources occurs; isotopic ratios 

sometimes are hard to determine in environmental samples 

due to analytical difficulties. It appears that use of δ15N to 

identify sources of nitrogen to ground waters is feasible 

only in relatively simple systems (e.g., low rainfall, single 

or dual sources, and minimal nitrogen transformations such 

as denitrification). Watershed- or ecosystem- level studies 

using this technique are semiquantitative at best. In most 

cases where δ15N can be applied reliably, the sources of 

nitrogen are obvious or can readily be estimated by other, 

less expensive means (Keeney, 1986). In one case, nitrogen 

isotopic data were of little help in determining the relative 

importance of cyanide and explosive residues as nitrogen 

sources in surface ponds at three Nevada gold mine sites. 

The data do, however, provide strong evidence for natural 

attenuation of non-cyanide nitrogen species by dispersal 

of ammonia gas or by dispersal of N2O or N2 gas produced 

by denitrification of nitrate (Johnson et al., 2000). 

Use of 1H Nuclear Magnetic Resonance for 

Identifying Sources of Nitrate 

Dissolved organic matter (DOM) originating from a certain 

source usually carries characteristic marks in its molecular 

structures that can be recognized by spectroscopic analysis 

such as nuclear magnetic resonance (NMR). Sources of 

water-born contaminants, such as nitrate, can be identified 

by recognition of the characteristics of DOM entrained in 

the water. Lu et al. (2004) analyzed DOM in ground water 

collected from a dairy/crop production area (Chino Basin, 

CA) using 1H NMR. Results showed that DOM derived from 

natural soil organic matter has a characteristic resonance 

at a chemical shift region of 4.0-4.3 ppm, while DOM derived from dairy wastes has a characteristic resonance at 

a lower chemical shift region of 3.2-3.6 ppm. These signature resonances were then used to distinguish the origins 

of nitrate in the ground water. It was found that disposal 

of dairy wastes on croplands was the primary source of 

nitrate contamination in ground water underlying the Chino 

Basin dairy area.

Denitrification Enzyme Activity

Denitrification enzyme activity (DEA) experiments can be 

conducted according to Tiedje (1994) using soils/sediments 

collected from the aquifer. It is recommended that these 

tests be conducted with caution due to potential disturbance 

to intact soil/sediment conditions during retrieval and processing. However, DEA has been shown to be strongly 

related to annual denitrification rates in temperate zone 

soils, and is useful for comparisons within a site as well 

as for the distribution of rates within soil horizons (Groffman and Tiedje, 1989; Schnabel et al., 1996; Flite et al., 

2001). The end product of denitrification is N2 gas, thus, 

it may be useful to analyze ground water for dissolved N2

gas at locations where soils/sediments are sampled for 

DEA experiments. This dissolved gas is a good indicator 

of denitrification when found in excess of concentrations 

expected relative to atmospheric trace gases such as Ar 

(Vogel et al., 1981; Böhlke and Denver, 1995).

Estimation of Denitrifying Capacity

Ground-water denitrification capacity may be evaluated 

through laboratory microcosm studies using aquifer sediments or through direct in-situ tracer tests in discrete locations within the aquifer. However, successful extrapolation 

of the results from these measurements across the site will 

depend on the degree of sampling frequency to capture 

spatial heterogeneity in both horizontal and vertical dimensions of the site with respect to hydrology, soil characteristics, and biogeochemical processes. It is recommended 

that both approaches be employed for a given site, since 

there is evidence that measured potentials of nitrate removal 
94

derived from laboratory microcosms may overestimate removal capacities in the field (Kellman, 2004). Korom et al. 

(2005) illustrated the use of in-situ mesocosm to determine 

denitrification rates in aquifer sediments. These authors 

incorporated information from tracer tests, δ15N data, and 

geochemical data to determine that 58% of the denitrification was caused by S(-I) in pyrite as the electron donor with 

the rest from ferrous iron and organic carbon.

To quantify ground water denitrification in discrete locations of riparian aquifers, Addy et al. (2002) modified and 

evaluated an in-situ method based on conservative tracers 

and 15N-enriched nitrate. Ground water was “pushed” (i.e., 

injected) into a mini-piezometer and then “pulled” (i.e., 

extracted) from the same mini-piezometer after an incubation period. This push–pull method was applied in replicate 

mini-piezometers at two Rhode Island riparian sites, one 

fresh water and one brackish water. Conservative tracer 

pretests were conducted to determine incubation periods,

ranging from 5 to 120 h, to optimize recovery of introduced

plumes. For nitrate push–pull tests, they used two conservative tracers, sulfur hexafluoride and bromide, to provide 

insight into plume recovery. The two conservative tracers 

behaved similarly.The dosing solutions were amended with 15N-enriched nitrate to quantify the mass of denitrification 

gases generated during the incubation period. The in-situ

push–pull method detected substantial denitrification rates 

(mean of 97 μg N kg-1 d-1) at a depth of 65 cm in a glacial 

outwash setting located within 10 m of the stream.At the 

brackish site, high rates of ground water denitrificationwere 

found in marsh locations (mean of 123 μg N kg-1 d-1) and 

minimal denitrification in soils fringing the marsh (mean of 

2 μg N kg-1 d-1) at a depth of 125 cm. The push–pull method 

can provide useful insights into spatial and temporal patterns of denitrification in riparian zones. The method is 

robust and results are not seriously affected by dilution or 

degassing from ground water to soil air. In conjunction with 

measurements of ground water flow, this method holds

promise for evaluating the influence of site and management factors on the ground-water nitrate removal capacity 

of riparian zones (Addy et al., 2002; Kellogg et al., 2005).

Attempts have been made to estimate the denitrifying removal capacity of an entire wetland (Maître et al., 2005). 

The authors developed a methodology that consisted of 

the following steps: 1) delineating the structure of aquifer 

formation in the riparian wetland; 2) mapping the spatial 

variation of the thickness of the different soil horizons 

constituting the aquifer formation; 3) measuring the DEA 

for each type of soil horizon; 4) grouping the soil horizons 

in specific soil-denitrifying classes; 5) mapping the spatial 

variation of the thickness of each soil-denitrifying class; 

6) mapping the spatial variation of the water table position; 7) combining the maps to calculate the volume of the 

fraction of each soil-denitrifying class that interacts with 

groundwater; and 8) calculating the denitrifying removal 

capacity of the whole site. 

Long-term Capacity

The long-term capacity of nitrate transformation within the 

aquifer will depend on the prevailing ground-water chemistry and the dynamics of biotic and abiotic processes 

controlling conversion of nitrate to other nitrogen species. 

If there are significant changes in ground-water chemistry 

over time, then the potential exists for reductions in the 

aquifer denitrifying capacity relative to nitrate loading from 

the contaminant plume. For situations in which nitrate 

transformation is controlled by biotic (microbial) processes, 

this may be due to reductions in degradable organic carbon 

(dissolved or associated with aquifer matrix) and nutrients 

needed to sustain the microbial community. For situations 

where abiotic components control nitrate transformation, 

e.g., reduced Fe-bearing minerals, changes to more oxidizing conditions or reductions in the supply of iron to sustain 

the mass of reactive minerals would likely result in nitrate 

plume expansion. Ultimately, the evolution of biogeochemical conditions in the subsurface will dictate the success of 

the MNA remedy. 

Since heterotrophic denitrification reactions require that the 

water is anaerobic (Eh < 350 mV) and that considerable 

organic carbon is present at concentrations greater than the 

nitrate concentration (Korom, 1992), significant widespread 

denitrification in aquifers are unlikely to occur. Based on 

water chemistry data, it was concluded that denitrification cannot be relied upon to decrease elevated nitrate 

concentrations in the modern, polluted, recharge waters 

of the English chalk (limestone) aquifer (Howard, 1985). 

On the other hand, situations exist where denitrification 

may be occurring in some shallow aquifers. For example, 

Egboka (1984) studied six ground water flow systems in 

Ontario, Canada under widely varying land use. Several 

of the ground waters exhibited low dissolved oxygen and 

Eh, but there was little consistency in the nitrate to chloride 

ratio. Nitrate concentrations at the sites that were reducing (low Eh) declined with distance from the nitrate source. 

Vidon and Hill (2004) observed effective nitrate removal 

by denitrification in riparian zones with hydric soils as well 

as in non-hydric riparian zones, and they concluded that 

a shallow water table is not always necessary for efficient 

nitrate removal by denitrification.

Chemolithoautotrophic denitrification may occur with pyrite 

as the electron donor. This is the proposed mechanism 

derived from a recent study based on the low dissolved 

organic carbon concentrations (< 1.5 mg L-1) that could not 

support heterotrophic denitrification, the common occurrence of disseminated pyrite in the aquifer, and the trend 

of increasing sulfate as ground water flowed from aerobic, 

unconfined conditions to anoxic, confined aquifer conditions 

(Beller et al., 2004). In this study, several independent 

lines of evidence suggested that microbial denitrification 

was naturally attenuating nitrate in a confined, O2-depleted 

region of a bedrock aquifer at a Lawrence Livermore National Laboratory site. The evidence included the following 

observations: (1) both nitrate and dissolved oxygen concentrations in ground water decreased dramatically as ground 

water flowed from unconfined to confined aquifer conditions, 

(2) stable isotope signatures (i.e., δ15N and δ18O) of ground 

water NO3

-

 indicated a trend of isotopic enrichment that was 

characteristic of denitrification, and (3) dissolved N2 gas, 

the product of denitrification, was highly elevated in nitrate-
95

depleted ground water in the confined region of the aquifer. 

At this site, long-term nitrate concentrations were relatively 

high and constant in recharge-area monitoring wells (typically 70 – 100 mg NO3

-

 L-1) and relatively low and constant 

in the downgradient confined region (typically < 0.1 – 3 mg 

NO3

-

 L-1), suggesting a balance between rates of nitrate 

loading and removal by denitrification. Autotrophic denitrification with pyrite (or other reduced Fe-S minerals) has 

been invoked to explain geochemical results of other field 

studies where low dissolved organic carbon concentrations 

could not explain apparent nitrate degradation in ground 

water (Kölle et al., 1985; Böttcher et al., 1990; Postma et al., 

1991; Robertson et al., 1996; Pauwels et al., 1998; Aravena 

and Robertson, 1998; Tesoriero et al., 2000). 

Tiered Analysis

Determination of the viability of nitrate remediation in ground 

water via monitored natural attenuation will depend upon 

proper assessment of contaminant loading to the aquifer 

and prevailing geochemistry and mineralogy within the 

contaminant plume and the down gradient zone prior to 

the point(s) of compliance. It is essential to understand 

the cause-and-effect relationship between loss of nitrate in 

ground water and the mechanisms responsible for its loss. 

The following tiered analysis structure for site characterization provides an approach to evaluate candidate sites and 

define the potential limitations of MNA as part of a remedy 

for ground-water cleanup.

Tier I. Site characterization under Tier I will involve demonstration that the plume is static or shrinking, has not 

reached compliance boundaries, and does not impact 

existing water supplies. Once this is established through 

ground-water characterization, evidence is collected to 

demonstrate nitrate reduction within the plume. Rapid 

movement of contaminants along preferred flow paths in 

the unsaturated and saturated zones can be induced by 

hydrologic events such as heavy rains (e.g., McCarthy et 

al., 1998; Camobreco et al., 1996). It will be important to 

determine that such hydrogeologic features do not result in 

contaminants bypassing zones where natural attenuation is 

occurring. An observed decrease in nitrate concentration 

in space and time that is attributable to a mass-removal 

process is anticipated if natural attenuation processes are 

active throughout the plume. Conditions that would support 

this initial screening would include evidence of reducing 

conditions (i.e., low EH) and abundant electron donors (e.g., 

elevated total organic matter or the presence of degradable 

organic co-contaminants). Identification of nitrate removal 

along ground-water flow paths provides justification for 

proceeding to Tier II characterization efforts.

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

of attenuation are determined. Estimates of a site attenuation rate(s) can be determined via several approaches: 

1) assessment of nitrate disappearance across well transects along the ground-water flow path(s), 2) assessment 

of in-situ rates of nitrate reduction through analysis of 

changes in the nitrogen stabile isotope fractionation along 

the ground-water flow path, and 3) ex-situ assessment of 

nitrate reduction rate(s) through the use of microcosm studies. In addition, time-series data may be collected at one or 

more monitoring points within the plume. This information 

will allow assessment of the relative timescale(s) for nitrate 

reduction relative to ground-water seepage velocities and 

determination of whether remediation objectives can be met 

within the required regulatory time frame. In addition, the 

mechanism(s) for attenuation need to be identified under 

this stage of site characterization (e.g., See site characterization approach in Postma et al., 1991.). This effort may require identification of the microbial agents involved in nitrate 

reduction and their distribution throughout the plume (see 

Site Characterization section in this chapter). Alternatively, 

sites at which an abiotic process dominates nitrate reduction 

may require characterization of aquifer solids to determine 

the presence of candidate mineral phases such as those 

discussed previously (see Equations 8.10 and 8.11). If a 

link is established between the apparent disappearance 

of nitrate and observed microbial or chemical processes 

active within the plume, then this provides justification for 

proceeding to Tier III characterization efforts.

Tier III. Once the nitrate reduction mechanism(s) have been 

identified for the site, the subsequent characterization effort 

under Tier III will involve determination of the capacity of 

the aquifer to sustain continued biological reduction. The 

impact of potential hydrologic changes, such as a shift in 

flow direction caused by the onset of pumping at nearby 

sites, needs to be determined. Since the active microbial 

community will reside within the boundaries of the nitrate 

plume, potential alterations in the predominant flow path 

may deliver nitrate to regions of the aquifer in which there 

is a reduced (or inactive) population of nitrate-reducing 

bacteria (NRB) to sustain reduction. A well-constrained 

ground-water flow model of the site relative to observed 

spatial distributions of nitrate and NRB density will assist 

in assessing the potential impact of such changes. For 

sites in which organic co-contaminants are identified as 

the predominant electron donor for nitrate reduction, it is 

recommended that an evaluation of the long-term supply 

of the electron donor be assessed. If engineered remedies 

for the treatment of co-contaminants are active or planned 

(e.g., source zone treatment or removal of co-contaminants), then it is expected that the impact these remedies 

have on the potential supply of electron donor will influence 

the sustainability of nitrate reduction. This may require 

evaluation of whether 1) the total organic carbon content 

and characteristics of treated ground water are sufficient to 

maintain NRB activity or 2) the available mass of reactive 

minerals within aquifer sediments is sufficient to sustain 

chemical reduction. Analysis of the capacity for sustained 

biological reduction could be conducted through laboratory 

microcosm studies, whereas assessment of the long-term 

capacity for chemical reduction may be achieved through 

assessment of the mass of reactive minerals within the 

plume flow path and their stability relative to anticipated 

changes in ground-water chemistry. 

For sites in which uncontaminated portions of the aquifer 

are aerobic (i.e., oxygen is present) or have naturally 
96

elevated concentrations of competing electron acceptors 

such as perchlorate, it is anticipated that evolution of the 

plume geochemistry may evolve to a condition that will not 

sustain microbial nitrate reduction. It is recommended that 

the potential impact of anticipated changes in ground-water 

chemistry on the rate or sustainability of nitrate reduction 

be assessed through microcosm studies. Ultimately, the 

ability to forecast the potential impacts of changes in aquifer 

chemistry or ground-water flow can be improved through 

the development of a reaction-transport model that includes 

site-specific parameterization of the microbial or chemical 

reduction process and site hydrogeology. If site-specific 

tests demonstrate sufficient capacity within the aquifer to 

sustain nitrate reduction, then the site characterization effort 

can progress to Tier IV. For cases where aquifer capacity 

is insufficient for plume reduction to required levels, then a 

determination of the benefits of contaminant source reduction or removal may be necessary.

Tier IV. Finally, under Tier IV a monitoring plan is established along with contingency plans in the event of MNA 

failure. It is recommended that the monitoring plan be 

designed to establish both continued plume stability and to 

identify changes in ground-water chemistry that may lead 

to a decrease in the rates or capacity for nitrate reduction. 

The specific chemical parameters to be monitored will 

include those identified under Tier III that may halt or slow 

down microbial nitrate reduction within the plume. For example, solution phase parameters that could alter nitrate 

reduction include the concentrations of dissolved oxygen or 

competing electron acceptors such as perchlorate. Similarly, a decrease in the concentration of electron donors 

such as natural organic matter, organic co-contaminants, 

or reactive minerals could slow or halt nitrate reduction 

via biological or abiotic processes. Changes in these parameters may occur prior to observed changes in nitrate 

concentrations and, thus, serve as monitoring triggers for 

potential MNA failure. In this instance, a contingency plan 

can be implemented that incorporates strategies to arrest 

possible plume expansion beyond compliance boundaries. 

Possible strategies to prevent plume expansion include 

pump and treat operations, installation of reactive barriers 

to enhance reduction capacity perpendicular to the direction 

of plume advance, or enhancement of natural attenuation 

processes within the aquifer through the injection of soluble 

reactive components (see Remedial Technologies section 

in this chapter).

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101

Perchlorate

Patrick V. Brady, Richard T. Wilkin, Robert G. Ford

Occurrence and Distribution

Ammonium perchlorate (NH4ClO4) has been used with 

powdered aluminum and various chemical binders as rocket 

fuel and in munitions since the mid-1940s. It powers the 

space shuttle and the U.S. nuclear missile arsenal. There 

are no ready substitutes. Potassium perchlorate is used in 

airbags in cars, in the production of leather, fabrics, coloring agents, fireworks (Wilkin et al., 2007), and elsewhere. 

Perchlorate in the environment is largely associated with the 

periodic replacement of perchlorate in rockets (ammonium 

perchlorate has a long, but not infinite, shelf-life) and its 

subsequent disposal. Naturally-occurring perchlorate from 

nitrate deposits in Chile has occasionally been released to 

the environment during mining to recover nitrate for use as 

fertilizer (e.g., Urbansky et al., 2001). Perchlorate has also 

been detected in sylvite from New Mexico, as well as in langbenite – a potassium sulfate mineral. The most common cocontaminants found at perchlorate-contaminated sites are 

nitrate and sulfate (ITRC, 2005). Atmospheric sources of 

perchlorate have also been implicated as a potential source 

of perchlorate in arid and semi-arid regions (Rajagopalan et 

al., 2006). Perchlorate is a powerful oxidant when heated, 

but the oxidation reaction is kinetically sluggish at low 

temperatures. The combination of low temperature stability 

and high temperature reactivity is what makes perchlorate 

so attractive for explosives and rocket fuel. 

Plume Characteristics

Because perchlorate is a large anion with a relatively low 

diffuse charge it is non-complexing, forms no insoluble 

minerals, and sorbs poorly to most solids. The solubility of 

perchlorate salts is typically in the range of 10 - 2000 g L-1. 

Although perchlorate sorption to soil materials is typically 

assumed to be negligible, there is some evidence that 

perchlorate might sorb to soil organics (e.g., Urbansky and 

Brown, 2003). Partitioning coefficients (i.e., Kd) describing 

sorption to geologic materials are usually found to equal 

zero. Few ion exchange resins or filter media effectively 

remove perchlorate from drinking water or wastewater. 

Similarly, perchlorate sorbs poorly to most geologic materials, suggesting that perchlorate plumes would likely move 

at roughly the same speed as ground water in the absence 

of biodegradation. Perchlorate-rich brines released from 

industrial applications can be sufficiently dense as to ‘sink’ 

in the subsurface analogous to dense non-aqueous phase 

liquids (Flowers and Hunt, 2000; Motzer, 2001). 

Remedial Technologies

A current review of available technologies for remediation 

of perchlorate contamination in ground water has been 

published by the Interstate Technology Regulatory Council 

(ITRC, 2005), which is available via the internet at http://

www.itrcweb.org/Documents/PERC-1.pdf. Available technologies include: 1) ex-situ, pump-and-treat approaches 

that extract perchlorate from recovered ground water or 

degrade perchlorate to innocuous compounds through 

biotic or abiotic chemical reduction, and 2) in-situ treatment approaches that degrade perchlorate within the 

plume via stimulation of native microbial communities or 

the emplacement of media that achieve biotic and abiotic 

chemical reduction of perchlorate, respectively. The success of these treatment approaches will be influenced by 

the level of knowledge of the ground-water flow regime as 

well as the geochemical characteristics within the plume. 

For technologies based on perchlorate extraction, the presence of common anions in ground water may compete for 

binding sites on ion exchange resins or other solid media 

used in the treatment process. For microbial perchlorate 

reduction (ex-situ or in-situ), geochemical parameters such 

as low pH and high total dissolved solids, as well as the 

presence of aerobic conditions or competing electron acceptors such as nitrate or sulfate may negatively impact 

the rates and extent of perchlorate reduction that can be 

achieved. It is important to consider these site-specific 

factors when selecting treatment technologies.

Regulatory Aspects

Introduction of a new analytical method in 1997 by the 

California Department of Health Services lowered the detection limit for perchlorate from 400 ppb to 4 ppb. Since 

that time relatively wide occurrences of perchlorate have 

been detected. As of 2001, forty-four states have perchlorate manufacturers or users (Motzer, 2001). There were 

24 sites on the National Priorities List with perchlorate 

contamination as of September 23, 2004 (USEPA, 2004) 

with various cleanup actions having been initiated at 12 

of these sites (http://www.epa.gov/swerffrr/documents/

perchlorate_site_summaries.htm). Perchlorate has been 

found in the water supplies of over 15 million people in 

California, Nevada and Arizona and in surface or ground 

water throughout the United States. 

Perchlorate is used to treat Grave’s Disease (hyperthyroidism) and in sufficient quantity competes with iodide in the 
102

thyroid gland. The perchlorate ion (ClO4

-

) has a charge 

and radius close to iodide and substitutes for iodide in the 

thyroid causing a decrease in hormonal output. Perchlorate 

is apparently not metabolized in the body, nor is it rapidly 

reduced to chloride in many soil settings. 

There are at present no state or federal drinking water 

standards for perchlorate, but several are being considered. With authority under the Safe Drinking Water Act 

(SDWA), in March 1998, the U.S. Environmental Protection Agency’s Office of Water formally added perchlorate 

to the drinking water contaminant candidate list (CCL) 

(Urbansky and Schock 1999). A National Primary Drinking Water Regulation (NPDWR) has not yet been promulgated for perchlorate, but monitoring will continue under 

the Unregulated Contaminants Monitoring Rule (UCMR). 

Several states have set action or advisory levels for perchlorate: Nevada – 18 ppb; Arizona – 14 ppb; California and 

Texas – 4 ppb; Maryland and Massachusetts – 1 ppb. No 

state governments have set enforceable standards. Current efforts are focused on identifying the specific health 

risks posed by perchlorate intake (National Research 

Council, 2005; http://www.nap.edu/catalog/11202.html). 

Perchlorate compounds do not have reportable quantities under the Comprehensive Environmental Response, 

Compensation, and Liability Act (CERCLA), as amended, 

but the USEPA Office of Solid Waste and Emergency 

Response has issued recent guidance with a Drinking 

Water Equivalent Level (DWEL) for actions administered 

under the CERCLA of 24.5 μg L-1 based on an adopted 

reference dose (RfD) for perchlorate of 0.0007 mg/kg-day 

(http://www.epa.gov/fedfac/pdf/perchlorate_guidance.pdf). 

Geochemistry and Attenuation Processes

Attenuation of perchlorate might be achieved through 

reduction ultimately to chloride. Specialized microorganisms able to reduce perchlorate to chloride have been 

identified (Coates and Achenbach, 2004; Achenbach et 

al., 2001; Coates et al., 1999) that are widespread in the 

environment. All of the identified perchlorate reducers 

are facultatively anaerobic or microaerophilic (Coates et 

al., 1999). Perchlorate-reducing bacteria use as electron 

donors: hydrogen, organic acids and alcohols, aromatic 

hydrocarbons, reduced humic substances, hexoses, ferrous 

minerals and hydrogen sulfide (Chaudhuri et al., 2002). 

In situ perchlorate biodegradation methods typically use 

electron donors such as vegetable oils, organic acids, 

alcohols, or sugars. Anaerobic conditions are required for 

biodegradation of perchlorate to proceed. Metabolism of 

electron donors leads to the consumption of free oxygen 

and nitrate, at which point perchlorate reduction proceeds. 

Molybdate is apparently important for perchlorate reduction; 

nitrate occasionally is seen to inhibit perchlorate reduction 

(Chaudhuri et al., 2002).

Dozens of strains of perchlorate-reducing bacteria have 

been identified in the environment, their ubiquity, substrate 

specificity, and wide distribution suggesting both a 

natural source of perchlorate and the potential for natural 

attenuation of perchlorate. Perchlorate reduction to chloride 

proceeds in the general sequence of perchlorate to chlorate 

to chlorite to hypochlorite to chloride:

ClO4

- → ClO3- → ClO2- → ClO- → Cl-

 + O2 (Eq. 9.1)

The first step of perchlorate reduction to chlorite is thought 

to be the rate-limiting step in the cascade. 

The Chilean Perchlorate Deposits

Two lines of evidence suggest that natural perchlorate fluxes 

into soils might be appreciable and that natural attenuation 

of very small amounts of perchlorate is perhaps common. 

The specificity and relative ubiquity of perchlorate reducing microorganisms imply the existence of an undetected, 

“shadow” flux of perchlorate (Urbansky and Schock, 1999). 

As noted by Urbansky and Schock (1999),

“If it can be shown that perchlorate is produced 

naturally in the environment and yet levels are very 

low, we must conclude that natural attenuation is 

responsible for the dichotomy.”

Presumably, the development of more powerful analytical 

methods will point to the existence or non-existence of 

such a flux. 

The occurrence of natural perchlorate in the nitrate deposits of the Atacama Desert in Chile points to the existence 

of a natural process of perchlorate formation – though it 

says little more. Perchlorate minerals occur along with 

other oxidized minerals rarely observed elsewhere including nitrates, iodates, chromates and dichromates. There 

are competing explanations for the origin(s) of the nitrate 

salts – which have received the most attention - including 

atmospheric formation (through electrochemical or photochemical means) followed by evaporative concentration 

(Bohlke et al., 1997; Dasgupta et al., 2005) or fixation by 

microorganisms (Ericksen, 1983). Ericksen (1983) suggested a photochemical oxidation of atmospheric Cl and O 

in the atmosphere or at ground level to form the perchlorate 

deposits, citing as evidence arguments that a similar reaction might take place in the stratosphere. The iodates in 

Chile might have come from sea spray (much of the iodine 

in the sea is iodate). The formation path for chromate is 

less clear as there are few sources of Cr near the area. 

The dichromate mineral lopezite – K2Cr2O7 – appears to 

be an alteration product of the chromate minerals (Ericksen, 1983). Note that the half cell reaction for reduction of 

perchlorate to chloride is:

ClO4

-

 + 8e- + 8H+ → Cl- + 4H2O Eo = 1.29V (Eq. 9.2)

which means that in acidic solutions perchlorate is in 

theory a stronger oxidant than oxygen, but not as strong 

as dichromate (Urbansky and Schock, 1999). In other 

words, a dichromate mineral once formed might be able to 

oxidatively form perchlorate from chloride. In the absence 

of a clearer understanding of how perchlorate came to form 

in the Atacama Desert, two poorly constrained potential 

fluxes present themselves:
103

1. An undefined atmospheric source of perchlorate – perhaps oxidation of chloride to chlorate 

by OH radicals, followed by further photochemical oxidation by H2O2, ozone or HO2 radical 

(e.g., Dasgupta et al., 2005),

2. An undefined soil source of perchlorate linked 

to oxidation of chloride by possibly transient 

dichromate sources or other unidentified soil 

processes.

Obviously the above is largely speculation in the absence of 

a better understanding of mechanisms. Presumably, better 

analytical methods will allow more precise measurement 

of actual perchlorate backgrounds and fluxes. 

Site Characterization

Site characterization, conceptual evaluation of long-term 

stability and capacity, and tiered analysis for perchlorate 

is likely be similar to that done for other contaminants that 

naturally attenuate through reduction by ambient microorganisms – namely chlorinated solvents. Anaerobic conditions (nitrate reducing or lower) are critical as are sufficient 

quantities of potential electron donors. In the absence of 

further research demonstrating otherwise, it is reasonable 

to assume the presence of perchlorate-reducing bacteria 

(Wu et al., 2001). A recent publication outlines an approach 

used to assess the limitations for natural biodegradation of 

perchlorate at a contaminated site (Kesterson et al., 2005). 

These authors collected data on the presence of perchlorate-reducing bacteria, availability of electron donors, and 

the concentrations of competing electron acceptors. While 

the presence of perchlorate-reducing bacteria in soil and 

water was determined throughout the site, perchlorate 

persisted in the subsurface due either to the lack of available electron donors and/or the presence of competing 

electron acceptors. This work illustrates the importance 

of collecting site specific data to explicitly demonstrate 

that natural attenuation of perchlorate by biodegradation is 

both active and of sufficient magnitude to achieve required 

mass/concentration reductions in ground water. A general 

summary is presented in Table 9.1 of the attenuation process anticipated to dominate perchlorate transport in ground 

water along with a synopsis of the types of information to 

be collected to support site characterization.

Determining Cl Ground-water Speciation

Determination of the potential extent of perchlorate degradation within a plume can be initially assessed through 

observation of the mass distribution of chlorine-bearing 

inorganic ions according to the sequential reaction scheme 

shown in Equation 9.1. ITRC (2005) provides a review 

of several published methods for determining perchlorate 

or the distribution of Cl-bearing inorganic anions in water 

samples, including those similar to published EPA methods 

(Table 9.2). Unlike methods based on analyte detection by 

conductivity, methods employing mass spectrometry are 

capable of confirming the chemical composition of unknown 

ions by their molecular weight (mass-to-charge ratio). This 

may be important for samples with complex matrices containing ions that may co-elute with perchlorate during ion 

chromatography separation. Preservation studies suggest 

that filtration and refrigeration (4°C) of samples is sufficient 

for laboratory analysis within reasonable holding times 

(Stetson et al., 2006; Wilkin et al., 2007). Comparison of 

results from these studies suggests that filtration with a pore 

size less than 0.45 μm may be warranted to eliminate the 

potential for microbial degradation of perchlorate during 

storage. Since perchlorate is a common component of 

many detergents used in field decontamination procedures 

(such as Alconox, Alcotabs, Liqui-Nox, and Neutrad), it is 

recommended that equipment rinseate blanks be collected 

during each sampling event when sampling equipment is 

reused (ITRC, 2005). Thorne (2004) developed a field 

colorimetric method that showed good agreement with EPA 

Method 314.0 over the range of 1–225 μg L-1 (slope = 1.11, 

R2 = 0.913) for well water and bioreactor effluent samples. 

The method is based on pre-concentration of perchlorate 

onto a solid-phase extraction cartridge that has been conditioned with a perchlorate-specific ion-pair reagent, followed 

by elution into an ion-pairing dye that is further treated for 

absorbance measurement at 640 nm using a standard 

portable spectrophotometer. No other published literature 

was available to further document the field performance of 

this method, but the general simplicity of the method suggests that it may be suitable as a field screening tool when 

there is an immediate need for analytical data to guide the 

placement of additional ground-water sampling locations 

during initial site characterization efforts. 

Table 9.1 Natural attenuation and mobilization pathways for perchlorate. 

Attenuation 

Processes

Mobilization 

Processes Characterization Approach

Biotic (microbial) or 

abiotic transformation 

of perchlorate to other 

chlorine species

Cessation of microbial 

reduction processes via 

changes in electron donor 

or nutrient supply within the 

ground-water flow path

Determination of the mass distribution of perchlorate and 

other Cl-bearing inorganic ions in ground water. Determination of water chemistry, microbial populations, and/or 

sediment mineralogy linked to perchlorate transformation 

along with spatial and temporal variability of required chemical/microbial components relative to perchlorate transport 

pathway(s).
104

Table 9.2 Published USEPA methods for determination 

of perchlorate and other Cl-bearing inorganic 

Ions in aqueous samples.

Method Name and Number Source

Determination of Perchlorate in 

Drinking Water by Ion Chromatography; 314.0

November 1999

http://www.epa.

gov/ogwdw/methods/pdfs/met314.

pdf

Determination of Perchlorate 

using Ion Chromatography with 

Chemical Suppression Conductivity Detection; 9058, Revision 0

November 2000

http://www.epa.

gov/epaoswer/

hazwaste/test/

pdfs/9058.pdf

Perchlorate in Water, Soils and 

Solid Wastes Using High Performance Liquid Chromatography/Electrospray Ionization/Mass 

Spectrometry; 6850, Revision 0

January 2007

http://www.epa.

gov/epaoswer/

hazwaste/test/

pdfs/6850.pdf

Perchlorate in Water, Soils and 

Solid Wastes Using Ion Chromatography/Electrospray Ionization/

Mass Spectrometry 

(IC/ESI/MS OR IC/ESI/MS/MS); 

6860, Revision 0

January 2007

http://www.epa.

gov/epaoswer/

hazwaste/test/

pdfs/6860.pdf

Determination of Perchlorate 

in Drinking Water Using Inline 

Column Concentration/Matrix 

Elimination Ion Chromatography 

with Suppressed Conductivity 

Detection; 314.1, Revision 1.0

EPA 815-R-05-

009, May 2005

http://www.epa.

gov/ogwdw/methods/pdfs/method_

314_1.pdf

Determination of Perchlorate in 

Drinking Water by Liquid Chromatography Electrospray Ionization Mass Spectrometry, 331.0, 

Rev. 1.0

EPA 815-R-05-

007, January 2005

http://www.epa.

gov/ogwdw/methods/pdfs/met331_

0.pdf

Determination of Perchlorate 

in Drinking Water by Ion Chromatography with Suppressed 

Conductivity and Electrospray 

Ionization Mass Spectrometry; 

332.0

EPA/600/R05/049, March 

2005

http://www.epa.

gov/nerlcwww/

m_332_0.pdf

Determination of Inorganic 

Anions by Ion Chromatography; 

9056A, Revision 1

November 2000

http://www.epa.

gov/epaoswer/

hazwaste/test/

pdfs/9056a.pdf

Stable Isotope Techniques

Stable isotopes of chlorine and oxygen within the perchlorate anion can be used to aid in differentiating natural and 

synthetic sources of contaminant as well as the degree to 

which perchlorate has been reduced by biotic processes. 

Specifically, determination of the distribution of the stable 

isotopes of oxygen (18O, 17O, 16O) can be used to determine 

the value of ∆17O (or 17O anomaly), defined by the following equations:

δ17O = (17Cl/16Cl)sample/(17Cl/16Cl)reference – 1 (Eq. 9.3)

δ18O = (18Cl/17Cl)sample/(18Cl/17Cl)reference – 1 (Eq. 9.4)

∆17O = [(1+ δ17O)/(1+ δ18O)0.525] – 1 (Eq. 9.5)

The value of ∆17O can be used to differentiate between natural sources of perchlorate (e.g., Atacama salt deposits and 

fertilizer products) versus synthetic perchlorate produced by 

electrolytic oxidation (Bao and Gu, 2004; Bohlke et al., 2005; 

Motzer et al., 2006). As discussed by Bao and Gu (2004), 

chemical reactions and biodegradation of perchlorate will 

not change the value of ∆17O of the original perchlorate, 

which can only occur within a ground-water plume through 

addition of another source of perchlorate with a different 

∆17O signature. In addition, the values of δ37Cl (defined 

as [(37Cl/35Cl)sample/(37Cl/35Cl)reference - 1] relative to Standard 

Mean Ocean Chloride; Long et al., 1993) from Atacama 

nitrate ore and from Chilean nitrate fertilizer products are 

the lowest reported for any common substance on Earth 

(Bohlke et al., 2005). These measurements have been 

used to determine the sources of perchlorate in groundwater plumes at several sites of contamination (Bohlke et 

al., 2005).

Changes in the fractionation of the chlorine stable isotope 

composition of perchlorate are significant during microbial 

reduction (Coleman et al., 2003; Sturchio et al., 2003), 

consistent with preferential reduction of perchlorate containing the lighter chlorine isotope (35Cl). Coleman et al. 

(2003) observed chlorine isotope fractionation on the order 

of -15‰ during perchlorate reduction by a microorganism 

isolated from a swine waste lagoon (Dechlorosoma suillum

strain PS). Sturchio et al. (2003) also assessed chlorine 

stable isotope fractionation during perchlorate reduction 

using Dechlorosoma suillum JPLRND isolated from ground 

water sampled in southern California. In addition, these 

authors used EPA Methods 300.0 and 314.0 to monitor the 

speciation of perchlorate, chlorate, chlorite, and chloride 

throughout incubation observing only minor intermittent 

concentrations of chlorate (<1.3 % of perchlorate concentration). The overall reaction was dominated by stoichiometric 

conversion of perchlorate to chloride. Based on these 

experiments, Sturchio et al. (2003) estimated that chlorine 

stable isotope fractionation during microbial perchlorate 

reduction may be detected at levels of biodegradation <2%, 

which may provide a more sensitive limit of detection than 

that obtained through direct measurements of changes in 

perchlorate and chloride concentrations within the groundwater plume. 
105

Stable isotope measurements for this contaminant have 

been made possible through the development of a highly 

selective bifunctional anion-exchange resin and sample processing method to extract and concentrate perchlorate from 

environmental samples (Gu et al., 2000; Gu et al., 2001). As 

reported by Bohlke et al. (2005), the collection of perchlorate 

onto the anion-exchange resin can be accomplished in the 

field with further processing conducted in the laboratory. 

However, the acquisition of a sufficient mass of perchlorate 

to support isotopic analyses for ground-water samples with 

1-10 μg L-1 perchlorate may require significant volumes of 

water (Duncan et al., 2005). Determination of the various 

isotopes of chlorine and oxygen is dependent on the conversion of perchlorate to various gaseous products that carry 

the specific isotope of interest. For example, Bohlke et al. 

(2005) describe the following procedures: 1) determination 

of δ18O following decomposition of perchlorate to carbon 

monoxide, 2) determination of ∆17O following decomposition 

of perchlorate to produce oxygen, and 3) determination of 

δ37Cl following decomposition of perchlorate to produce 

chloride that was subsequently reacted to produce methyl 

chloride (CH3Cl; also known as chloromethane). As these 

authors point out, there are currently no perchlorate isotopic 

reference materials available for the purpose of assessing 

performance for a given analytical facility.

Identifying Microbial Populations

Recent work has demonstrated the ubiquity of perchloratereducing bacteria at sites with perchlorate contamination in 

ground water (Waller et al., 2004). These microorganisms 

reduced perchlorate in microcosm studies with or without 

supplement of an electron donor (e.g., acetate, molasses), 

although degradation rates were greater for supplemented 

microcosms. In order to assess the degree of intrinsic 

bioreduction of perchlorate at a given site, information is 

needed to identify the conditions of the microbial process 

controlling perchlorate reduction throughout the perchlorate 

plume. Bender et al. (2004) have developed two metabolic 

primer sets designed to target the chlorite dismutase (cld) 

gene, which can be used to identify perchlorate-reducing 

bacteria in environmental samples by a qualitative polymerase chain reaction (PCR) approach. False positive 

identification of perchlorate-reducing bacteria may occur, 

since the approach targets only a single gene that may be 

present even though the microorganism lacks other genes 

necessary to carry out perchlorate metabolism (e.g., genes 

required for perchlorate reductase). However, subsequent 

analysis of cld-positive samples using perchlorate reductase 

probes could be carried out in order to eliminate false positive identifications (Bender et al., 2005). This approach may 

substitute for or supplement for enumeration of perchloratereducing bacteria using the Most Probable Number (MPN) 

procedure (Wu et al., 2001; Kesterson et al., 2005).

Long-term Capacity

Given reducing conditions and labile electron donor masses 

that substantially exceed the perchlorate source, a critical 

unknown for determining the long-term capacity for attenuation is likely to be the perchlorate reduction rate relative 

to the flux of perchlorate and required electron donors 

and essential nutrients to sustain the metabolic process. 

Reduction rates for perchlorate remain an area of ongoing research. There is no universally accepted means for 

estimating perchlorate reduction rates, although the use 

of stable isotopes of chlorine and oxygen may provide a 

means to estimate in-situ rates of reduction along a plume 

flow path. Since there are potentially large uncertainties 

associated with extrapolating degradation rates determined 

from laboratory microcosm studies to the field setting, 

multiple lines of evidence based on field observations will 

likely provide the most reliable means for assessing the 

long-term capacity of the aquifer. In general, a default approach might then be to: 1) establish that conditions are 

reducing; 2) demonstrate that there are more than sufficient 

electron donors available for perchlorate reduction; 3) fit 

plume contours over time to estimate a biodegradation rate 

(possibly supported by microcosm studies demonstrating 

that perchlorate reduction is possible under the conditions 

prevailing in the field); and 4) use the derived model to assess long-term capacity for sustaining biological reduction 

of perchlorate.

Tiered Analysis

Determination of the viability of perchlorate remediation in 

ground water via monitored natural attenuation will depend 

upon proper assessment of contaminant loading to the 

aquifer and prevailing geochemistry and mineralogy within 

the contaminant plume and the down gradient zone prior to 

the point(s) of compliance. It is essential to understand the 

cause-and-effect relationship between loss of perchlorate in 

ground water and the mechanisms responsible for its loss. 

The following tiered analysis structure for site characterization provides an approach to evaluate candidate sites and 

define the potential limitations of MNA as part of a remedy 

for ground-water cleanup.

Tier I. Site characterization under Tier I will involve 

demonstration that the plume is static or shrinking, has 

not reached compliance boundaries, and does not impact 

existing water supplies. Once this is established through 

ground-water characterization, evidence is collected to 

demonstrate perchlorate reduction within the plume. Rapid 

movement of contaminants along preferred flow paths in 

the unsaturated and saturated zones can be induced by 

hydrologic events such as heavy rains (e.g., McCarthy et 

al., 1998; Camobreco et al., 1996). It will be important to 

determine that such hydrogeologic features do not result in 

contaminants bypassing zones where natural attenuation 

is occurring. In addition, for sites in which the suspected 

source term may have consisted of a perchlorate brine, it is 

recommended that the hydrogeologic characterization of the 

site establish whether perched ‘DNAPL-like’ plumes exist 

that could influence both the timeframe for remediation and 

the required attenuation capacity within the aquifer. If natural 

attenuation processes are active throughout the plume, then 

an observed decrease in perchlorate concentration in space 

and time that is attributable to a mass-removal process 

is anticipated. Conditions that would support this initial 

screening would include evidence of reducing conditions 
106

(i.e., low EH), abundant electron donors (e.g., elevated total 

organic matter or the presence of degradable organic cocontaminants), and the apparent availability of essential 

nutrients such as molybdate. Identification of perchlorate 

removal along ground-water flow paths provides justification 

for proceeding to Tier II characterization efforts.

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

attenuation are determined. Estimates of a site attenuation 

rate(s) can be determined via several approaches: 1) assessment of perchlorate disappearance across well transects along the ground-water flow path(s), 2) assessment 

of in-situ rates of perchlorate reduction through analysis of 

changes in the chlorine stabile isotope fractionation along 

the ground-water flow path, and 3) ex-situ assessment of 

perchlorate reduction rate(s) through the use of microcosm 

studies. In addition, time-series data may be collected at 

one or more monitoring points within the plume. This information will allow assessment of the relative timescale(s) for 

perchlorate reduction relative to ground-water seepage velocities and determination of whether remediation objectives 

can be met within the required regulatory time frame. In 

addition, the mechanism(s) for attenuation need to be identified under this stage of site characterization. This effort 

may require identification of the microbial agents involved 

in perchlorate reduction and their distribution throughout the 

plume (see Site Characterization section in this chapter). If 

a link is established between the apparent disappearance of 

perchlorate and observed microbial processes active within 

the plume, then this provides justification for proceeding to 

Tier III characterization efforts.

Tier III. Once the perchlorate reduction mechanism(s) 

have been identified for the site, the subsequent characterization effort under Tier III will involve determination of 

the capacity of the aquifer to sustain continued biological 

reduction. The impact of potential hydrologic changes, such 

as a shift in flow direction caused by the onset of pumping 

at nearby sites, needs to be determined. Since the active 

microbial community will reside within the boundaries of the 

perchlorate plume, potential alterations in the predominant 

flow path may deliver perchlorate to regions of the aquifer in which there is a reduced (or inactive) population of 

perchlorate-reducing bacteria to sustain reduction. A wellconstrained ground-water flow model of the site relative to 

observed spatial distributions of perchlorate and density 

perchlorate-reducing bacteria will assist in assessing the 

potential impact of such changes. For sites in which organic co-contaminants are identified as the predominant 

electron donor for perchlorate reduction, it is recommended 

that an evaluation of the long-term supply of the electron 

donor be assessed. If engineered remedies for the treatment of co-contaminants are active or planned (e.g., 

source zone treatment or removal of co-contaminants), 

then it is expected that the impact these remedies have 

on the potential supply of electron donor will influence the 

sustainability of perchlorate reduction. This may require 

evaluation of whether the total organic carbon content and 

characteristics of treated ground water are sufficient to 

maintain the activity of perchlorate reducing bactera. This 

analysis could be conducted through laboratory microcosm 

studies. For sites in which uncontaminated portions of the 

aquifer are aerobic (i.e., oxygen is present) or have naturally 

elevated concentrations of competing electron acceptors 

such as nitrate, it is anticipated that evolution of the plume 

geochemistry may evolve to a condition that will not sustain 

microbial perchlorate reduction. It is recommended that the 

potential impact of anticipated changes in ground-water 

chemistry on the rate or sustainability of perchlorate reduction be assessed through microcosm studies. Ultimately, 

the ability to forecast the potential impacts of changes in 

aquifer chemistry or ground-water flow can be improved 

through the development of a reaction-transport model 

that includes site-specific parameterization of the microbial 

reduction process and site hydrogeology. If site-specific 

tests demonstrate sufficient capacity within the aquifer to 

sustain perchlorate reduction, then the site characterization effort can progress to Tier IV. For cases where aquifer 

capacity is insufficient for plume reduction to required levels, 

then a determination of the benefits of contaminant source 

reduction or removal may be necessary.

Tier IV. Finally, under Tier IV a monitoring plan is established along with contingency plans in the event of MNA 

failure. It is recommended that the monitoring plan be 

designed to establish both continued plume stability and to 

identify changes in ground-water chemistry that may lead 

to a decrease in the rates or capacity for perchlorate reduction. The specific chemical parameters to be monitored will 

include those identified under Tier III that may halt or slow 

down microbial perchlorate reduction within the plume. For 

example, solution phase parameters that could alter perchlorate reduction include the concentrations of dissolved 

oxygen or competing electron acceptors such as nitrate. 

Similarly, a decrease in the concentration of electron donors 

such as natural organic matter or organic co-contaminants 

could slow or halt perchlorate biodegradation. Changes in 

these parameters may occur prior to observed changes in 

perchlorate concentrations and, thus, serve as monitoring 

triggers for potential MNA failure. In this instance, a contingency plan can be implemented that incorporates strategies 

to arrest possible plume expansion beyond compliance 

boundaries. Possible strategies to prevent plume expansion 

include pump and treat operations, installation of reactive 

barriers to enhance uptake/degradation capacity perpendicular to the direction of plume advance, or enhancement 

of natural attenuation processes within the aquifer through 

the injection of soluble reactive components (see Remedial 

Technologies section in this chapter).

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October 2007

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