United States
Environmental Protection
Agency
Office of
Research and Development
Washington, DC 20460
EPA/625/4-91/026
November 1991
Technology Transfer
EPA Seminar Publication
Site Characterization for
Subsurface Remediation
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EPA/625/4-91/026
November 1991
Seminar Publication
Site Characterization for Subsurface Remediation
Center for Environmental Research Information
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, OH 45268
Printed on recycled paper
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Notice
This document has been reviewed in accordance with U.S. Environmental Protection
Agency policy and approved for publication. Mention of trade names or commercial
products does not constitute endorsement or recommendation for use.
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Preface
Site characterization of contaminated sites has become an increasingly complex
process as a result of rapid developments in (1) methods for observing the physical,
chemical, and biological characteristics of the subsurface, and (2) methods for remediation
of soil and ground water. Consideration of the possible methods that may be used to clean
up contaminated soils or ground water early in the site characterization process can ensure
that data collected are appropriate and possibly reduce the time it takes to initiate clean-up
efforts.
This seminar publication provides a comprehensive approach to site characterization
for subsurface remediation. Chapter 1 describes a methodology for integrating site
characterization with subsurface remediation. This introductory chapter of the handbook
also provides a guide for quickly and efficiently accessing information in the rest of the
document for specific remediation applications through the use of summary tables,
checklists, figures, and flow charts.
The rest of the handbook is divided into three parts. Part I covers methods for
subsurface characterization, Part II covers physical and chemical processes in the subsur-
face that relate to the selection of remediation methods, and Part III covers methods for soil
and ground-water remediation.
In Part I, Chapter 2 provides an overview of the site characterization process. The next
four chapters cover physical aspects of site characterization: geologic and hydrogeologic
aspects (Chapter 3), characterization of water movement in the unsaturated zone (Chapter
4), characterization of the vadose zone (Chapter 5), and characterization of water move-
ment in saturated fractured media (Chapter 6). The remaining three chapters in Part I cover
geochemical aspects of site characterization: basic analytical and statistical concepts
(Chapter 7), the geochemical variability of the natural and contaminated subsurface
(Chapter 8), and geochemical sampling of soil and ground water (Chapter 9).
Part II contains three chapters on physiochemical processes affecting the transport of
major types of contaminants: organics in liquid and solid phases in the subsurface
(Chapter 10), organic volatilization and gas-phase transport (Chapter 11), and inorganic
contaminants (Chapter 12). Chapter 13 focuses on abiotic and microbiological degradation
and transformation processes in the subsurface.
Part III contains three chapters on remediation. Chapter 14 outlines basis approaches
to remediation of contaminated soil and ground water. The concluding chapters provide
more detailed information on specific techniques for cleaning up contaminated soil
(Chapter 15) and ground water (Chapter 16).
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Contents
Notice ii
Preface iii
Acknowledgments x
Chapter 1 Integrating Site Characterization with Subsurface Remediation 1
1.1 Approach for Integration of Site Characterization with Subsurface Remediation 1
1.2 Subsurface Site Characterization for Remediation Technology Selection 1
1.3 Site Reconnaissance 3
Part I: Methods for Subsurface Characterization
Chapter 2 Site Characterization 13
2.1 Introduction 13
2.2 Flow System Characterization 13
2.3 Contamination Characterization 16
2.4 Techniques for Characterization 17
2.5 Analysis of Data 18
2.6 References 20
Chapter 3 Geologic Aspects of Site Remediation 23
3.1 Stratigraphy 23
3.2 Lithology 24
3.3 Structural Geology 27
3.4 Hydrogeology 27
3.5 Hydrogeologic Investigations 28
3.5.1 Geophysical Techniques 28
3.5.2 Example - Hyde Park Landfill 32
3.6 References 38
Chapter 4 Characterization of Water Movement in the Saturated Zone 39
4.1 Review of Concepts 39
4.2 Field Techniques 41
4.2.1 Drilling Techniques 41
4.2.2 Methods to Measure Hydraulic Head 45
4.2.3 Methods to Determine Aquifer Properties 48
4.3 Analysis of Data 51
4.4 Remedial Actions 51
4.5 Example - Conservation Chemical Company Site 52
4.6 References 55
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Contents (continued)
Chapter 5 Characterization of the Vadose Zone 59
5.1 Review of Concepts 59
5.2 Field Techniques 61
5.2.1 Precipitation and Infiltration 61
5.2.2 Evaporation and Evapotranspiration 61
5.2.3 Moisture Content and Moisture Characteristics Curves 61
5.2.4 Vadose-Zone Hydraulic Conductivity 64
5.2.5 Soil Gas Analysis 64
5.3 Analysis of Data 66
5.4 Remedial Actions 66
5.5 Example-Pepper's Steel Site 68
5.6 References 69
Chapter 6 Characterization of Water Movement in Saturated Fractured Media 73
6.1 Review of Concepts 73
6.2 Field Techniques 74
6.2.1 Fracture Trace Analysis 74
6.2.2 Coring 75
6.2.3 Aquifer Tests 75
6.2.4 Tracer Tests 76
6.2.5 Geophysical Tools 76
6.2.6 Borehole Flowmeters 76
6.3 Analysis of Data 76
6.4 Remedial Actions 77
6.5 Example-Marion County, Florida 77
6.6 References 81
Chapter 7 Geochemical Characterization of the Subsurface:
Basic Analytical and Statistical Concepts 83
7.1 Data Measurement Reliability 83
7.1.1 Deterministic versus Random Geochemical Data 83
7.1.2 Data Representativeness 83
7.1.3 Measurement Bias, Precision, and Accuracy 84
7.1.4 Sources of Error 84
7.2 Analytical and QA/QC Concepts 87
7.2.1 Instrumentation and Analytical Methods 88
7.2.2 Limit of Detection 88
7.2.3 Types of Samples 89
7.3 Statistical Techniques 91
7.3.1 Statistical Approaches to Geochemical Variability 91
7.3.2 Geostatistics 91
7.4 Interpretation of Geochemical and Water Chemistry Data 92
7.4.1 Analysis of Censored Data 93
7.4.2 Contaminant Levels versus Background Conditions 93
7.5 References 98
Chapter 8 Geochemical Variability of the Natural and
Contaminated Subsurface Environment 103
8.1 Overview of Subsurface Geochemistry 103
8.1.1 Geochemical Processes 103
8.1.2 Environmental Parameters 103
VI
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Contents (continued)
8.1.3 The Vadose and Saturated Zones 106
8.2 Background Levels and Behavior of Chemical Constituents 107
8.3 Spatial Variability 108
8.3.1 Scale 108
8.3.2 Physical Gradients 108
8.3.3 Chemical Gradients 109
8.4 Temporal Variability 110
8.5 References 117
Chapter 9 Geochemical Sampling of Subsurface Solids and Ground Water 123
9.1 General Considerations 123
9.1.1 Types of Monitoring 123
9.1.2 Sampling Protocol 123
9.1.3 Sampling Location 125
9.1.4 Sampling Frequency 127
9.1.5 Sample Type and Size 128
9.1.6 Vadose versus Saturated Zone 129
9.2 Sampling Subsurface Solids and Vadose Zone Water 129
9.2.1 Analyte Selection 129
9.2.2 Sampling Devices and Techniques 129
9.3 Sampling Ground Water 130
9.3.1 Analyte Selection 131
9.3.2 Well Development 132
9.3.3 Purging 136
9.3.4 Well Construction and Sampling Devices 136
9.4 References 148
Part II: Physical and Chemical Processes in the Subsurface
Chapter 10 Physicochemical Processes: Organic Contaminants 155
10.1 Overview of Physicochemical Processes 155
10.2 Dissolution of Nonaqueous Phase Liquids 156
10.3 Sorption Phenomena 157
10.3.1 Adsorption Isotherms 157
10.3.2 Determining Retardation Factors Using foc and Koc 159
10.3.3 Determining Retardation Factors Using Batch Tests 161
10.3.4 Determining Retardation Factors Using Column Tests 162
10.3.5 Determining Retardation Factors From Field Data 163
10.3.6 Comparison of Methods for Estimation of Retardation 163
10.3.7 Applicability and Limitations of Linear Partitioning and Retardation 164
10.4 lonization and Cosolvation 164
10.5 Expressions for Other Chemical Processes 165
10.6 References 167
Chapter 11 Physicochemical Processes: Volatilization and Gas-Phase Transport 169
11.1 Volatilization 170
11.2 Gas-Phase Transport 170
11.2.1 Diffusion 170
11.2.2 Gas Phase Retardation 171
11.2.3 Processes Affecting Gas-Phase Transport 172
vn
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Contents (continued)
11.3 Vapor Extraction 175
11.4 References 178
Chapter 12 Physicochemical Processes: Inorganic Contaminants 179
12.1 Chemical Processes 180
12.1.1 Speciation 180
12.1.2 Dissolution/Precipitation 180
12.1.3 Oxidation/Reduction 181
12.1.4 Adsorption/Ion Exchange 182
12.2 Particle Transport 185
12.3 Organic-Inorganic Interactions 186
12.4 Computational Tools 189
12.4.1 Mass Balance 189
12.4.2 Chemical Speciation 189
12.4.3 Mass Transfer 189
12.4.4 Multicomponent Transport 189
12.5 References 190
Chapter 13 Characterization of Subsurface Degradation Processes 193
13.1 Abiotic Transformation Reactions 193
13.1.1 Hydrolysis 193
13.1.2 Substitution 194
13.1.3 Elimination 194
13.1.4 Oxidation-Reduction 194
13.2 Microbiological Transformations in the Subsurface 194
13.2.1 Microbial Ecology of the Subsurface 195
13.2.2 Relationship of Environmental Factors to Biodegradation 196
13.2.3 Microbial Metabolism 196
13.2.4 Biological Reaction Kinetics 197
13.3 Bioremediation of Organic Compounds in the Subsurface 198
13.3.1 General Considerations 198
13.3.2 Compounds Appropriate to Consider for Bioremediation 198
13.4 References 199
Part III: Soil and Ground-Water Remediation
Chapter 14 Soil and Ground-Water Remediation: Basic Approaches 203
14.1 Conceptual Approach to Soil and Ground-Water Remediation 203
14.2 Methodology 205
14.2.1 Site Characterization 206
14.2.2 Assessment of Problem 206
14.2.3 Treatment Approaches 209
14.2.4 Monitoring Program 210
14.3 Selection of Treatment Methods 210
14.3.1 Utility of Mathematical Models 210
14.3.2 Treatability Studies 210
14.3.3 Treatment Trains 211
14.4 Measurement and Interpretation of Treatment Effectiveness 211
14.5 References 212
Vlll
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Contents (continued)
Chapter 15 Remediation Techniques for Contaminated Soils 215
15.1 In Situ versus Prepared Bed Soil Remediation 215
15.2 In Situ Techniques 215
15.2.1 Soil Vacuum Extraction (SVE) 215
15.2.2 Bioremediation 229
15.2.3 Immobilization 232
15.2.4 Contaminant Mobilization 236
15.3 Prepared Bed Reactors 236
15.4 References 238
Chapter 16 Aquifer Restoration 243
16.1 Product Removal 243
16.2 Pump-and-Treat Remediation 244
16.3 Biorestoration 248
16.3.1 Example of the Use of Bioremediation: A Case Study 252
16.3.2 Advantages and Limitations in the Use of In Situ Bioremediation 257
16.4 References 258
IX
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Acknowledgments
This publication is based on the content of a series of U.S. Environmental Protection
Agency (EPA) technology transfer seminars that were conducted in all ten EPA Regions,
October 1989 through February 1990. This project was funded by the Office of Solid
Waste and Emergency Response (OSWER) and the Office of Research and Development
(ORD) to assist regulators and technical specialists in selecting the most appropriate
remediation technologies for contaminated soils and ground water at Superfund sites.
Seminar development was the responsibility of ORD staff in the Center for Environmental
Research Information (CERI), Cincinnati, OH, and the Robert S. Kerr Environmental
Research Laboratory (RSKERL), Ada, OK. Dominic DiGiulio, RSKERL, provided techni-
cal direction for seminar development and publication review. Marion R. Scalf, RSKERL,
and Carol Grove, CERI, were project managers. Seminars were held in October 1989
(Chicago, IL; Kansas City, MO; Denver, CO; and Dallas, TX); November 1989 (Lowell,
MA, and New York, NY); January, 1990 (Atlanta, GA, and Philalelphia, PA); and
February, 1990 (Seattle, WA, and San Francisco, CA).
Principal participants in the project include:
•Michael Barcelona, Institute for Water Science, Western Michigan University,
Kalamazoo, MI
•J. Russell Boulding, Eastern Research Group, Inc., Arlington, MA
•William Fish, Department of Environmental Science and Engineering, Oregon
Graduate Institute of Science and Technology
•J. Michael Henson, RMT Engineering and Environmental Management Services,
Greenville, SC
•Richard Johnson, Department of Environmental Science and Engineering, Oregon
Graduate Institute of Science and Technology
•James Mercer, GeoTrans, Inc., Sterling, VA
•Carl Palmer, Department of Environmental Science and Engineering, Oregon
Graduate Institute of Science and Technology
•Judith Sims, Utah Water Research Laboratory, Utah State University, Logan, UT
•Ronald Sims, Department of Civil and Environmental Engineering, Utah State
University, Logan, UT
•Charles Spalding, GeoTrans, Inc., Sterling, VA
Eastern Research Group, Inc., Arlington, MA, provided technical, editorial, and
production support for the project under Contract 68-C8-0014. Russell Boulding contrib-
uted as author, editor, and reviewer of the document Trisha Hasch provided seminar
coordination; and Karen Ellzey, Susan Richmond, Heidi Schultz, and Denise Short
provided editorial and production support.
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Chapter 1
Integrating Site Characterization with Subsurface Remediation
Ronald C. Sims and Judith L. Sims
This handbook on site characterization for subsurface
remediation emphasizes processes and concepts (Parts I and
II), characterization tools and analyses of data (Part I), and
remediation decisions (Part III). Chapter 1 relates subsurface
site characterization activities to the selection of subsurface
remediation technologies. Chapters 2 through 16 each address
a specific aspect of site characterization or remediation tech-
nology (e.g., geologic aspects, saturated zone, unsaturated
zone, remediation techniques for contaminated soils).
1.1 Approach for Integration of Site
Characterization with Subsurface
Remediation
Chapter 1 integrates the information presented in Chap-
ters 2 through 16 so that the reader is guided through the
Handbook and may access necessary interdisciplinary infor-
mation quickly and efficiently for specific remediation appli-
cations. The tables, checklists, figures, and flow charts in this
chapter synthesize relevant terms, parameters, and concepts
relating site characterization to specific subsurface remedia-
tion techniques. Using this information to select subsurface
treatment technologies requires specific information that is
interdisciplinary, thereby cutting across areas of specializa-
tion, i.e., chapters. Therefore, this chapter not only provides
an index to the Handbook, but also provides comments and
guidance about the relationship between characterization pa-
rameters and technology selection.
This chapter also discusses the importance of understand-
ing the surface physical layout of a site, including cultural
features and industrial structures (e.g., buildings, lots, produc-
tion units) and the evaluation of historical records of produc-
tion and waste management within the context of site
characterization for subsurface remediation. Activities such
as making site visits and obtaining historical records of site
and waste management are an integral part of site character-
ization. Information from these activities, which can provide
valuable insights concerning limitations as well as applica-
tions of remediation technologies at field scale, is referred to
collectively as site reconnaissance information.
1.2 Subsurface Site Characterization for
Remediation Technology Selection
A methodology for integrating site characterization with
subsurface remediation is shown in Figure 1-1. The develop-
ment of information for a specific site progresses from charac-
terization through monitoring (left to right as illustrated across
Figure 1-1). The figure presents characterization needs in
terms of waste interaction with unsaturated soil in the vadose
zone or sediment or aquifer material in the saturated zone as
influenced by site factors such as climate, topography, surface
slope, etc. Information from site characterization is used to
formulate, in qualitative and quantitative terms, the problem(s)
in terms of pathways of migration, escape, and/or exposure at
a contaminated site (problem assessment). This information is
used for subsurface treatment technique evaluation, elimina-
tion of unsuitable technique(s), and selection of an appropri-
ate treatment (train). Monitoring provides feedback on rate
and extent of remediation at field scale as well as information
for modification of site management. Sections 14.1 and 14.2
present this methodology in more detail.
Table 1-1 lists specific aspects of each step of the meth-
odology, presents relevant concepts, and indicates references
in the Handbook for additional information on each step of the
methodology. Specific characterization parameters are related
to problem assessment, treatment, and monitoring. For ex-
ample, the distribution coefficient, Kd, will allow evaluation
of the problem at a site with regard to migration. If soil
flushing is selected as a treatment technique, it may be moni-
tored effectively through pore-liquid phase sampling. Infor-
mation on each aspect can be found in the sections in the
Handbook listed under Text Reference (Section) on the table.
Subsurface-based waste characterization information needs
are summarized in Table 1-2. Potential impacts of waste on
ground-water, vadose-zone, atmosphere, and surface-water
resources depend upon properties of the waste chemicals and
properties of the affected matrix. Information on these proper-
ties is necessary to adequately assess the problem at a specific
site, as described above. Table 1 -2 presents individual param-
eters and text references for describing those parameters.
Figure 1-2 illustrates problem assessment in terms of
compartments as well as pathways of migration for chemical
migration, escape, and/or exposure. A mass balance concep-
tual approach to the subsurface identifies chemicals that will
(1) migrate upward (volatilization); (2) migrate downward
(leaching, pure product); (3) migrate laterally (aqueous plume
and pure product); and (4) remain in place as persistent
chemicals. A nonaqueous phase liquid (NAPL) may be fur-
ther classified as a light NAPL (LNAPL) if the density of the
1
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Characterization
Distribution
Reaction
Migration/Escape
Exposure
Problem Assessment
Treatment (train)
Monitoring
Figure 1-1. Methodology for integrating site characterization with subsurface remediation.
Table 1-1. Methodology for Relating Site Characterization to Subsurface Remediation
Character-
ization
Distribution
K<
«„
K.
species1'
Degradation
chemical
biological
Transport
advection
diffusion
Sampling
physical
environment
aqueous
environment
Modeling
vadose zone
saturated zone
Problem
Assessment
persistence
migration
loss in air
residential
saturation
phase(s) of
occurrence
rate/extent
intermediates
rate/extent
intermediates
extent/rate of
escape/
exposure
slow release
extent of
contamination
extent of
contamination
identify problem
identify problem
Treatment
biodegradability
soil flushing
vacuum extraction
soil flushing
biodegradation
immobilization
flushing, volatilizing
chemical destruction
and detoxification
biological destruction
and detoxification
containment, removal
destruction
containment, removal
destruction
accurate evaluation
of any technology
accurate evaluation
of any technology
for any treatment
for any treatment
Moni-
toring3
c
1
9
1
c
c,l,g
c.l
c,l
1.9
c,l,9
c
1
c. 1, 9
c,l,9
Text Reference (Section)
(10.3. 1) (12. 1.4) (14) (15.2.3) (16.2)
(10.3.2) (12.2) (14. 1) (14.2) (14.3) (15.2.4) (16.2)
(5.2.5) (11) (14.1) (14) (15.2.1)
(102.) (12.3) (14) (15.2.4) (16.2)
(13.3) (14) (15.2.2) (16.3)
(8. 1) (12. 1. 1) (12.4) (15.2.3) (15.2.4)
(12. 1) (13. 1) (14) (15. 1) (15.2.3)
(13.2) (13.3) (14) (15. 1) (15.2.2) (15.3) (16.3)
(2.2) (3) (4. 1) (5.2) (6. 1) (10. 1) (1 1.2) (12.2) (12.4.4)
(14) (15) (16)
(8.1) (8.3) (11.2.1) (12.4.3) (14) (16.2)
(2.5) (3. 1) (3.2) (3.3) (6.2) (7.3) (9.2)
(2.5) (3.4) (4.2) (5.2) (7.3) (8. 1) (8.3) (9.2) (9.3)
(5.3) (10. 1) (10.2) (10.3. 1) (10.3.2) (11.2) (12.4)
(13.2.4) (14.2. 1) (14.2.2) (14.3. 1)
(2.2) (3.5.2) (4.3) (4.5) (6.3) (10. 1) (10.2) (10.3. 1)
(10.3.2) (12.4) (13.2.4) (14.2. 1) (14.2.2) (16.3)
"(c) = core material; (1)
'Species are determined
= pore liquid phase; (g) = gas phase.
primarily for inorganics (metals) and affect metal phase (aqueous, solid, gas).
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liquid is less than water, or as a dense NAPL (DNAPL) if the
density of the liquid is greater than water. Additional informa-
tion on the compartments comprising the subsurface is pre-
sented in Section 14.1. Figure 1-2 also indicates references for
additional information on each topic.
Subsurface remediation techniques that may be evaluated
based upon site characterization and problem assessment, as
outlined above, are summarized in Table 1-3 and presented
for each technology-and-environment combination in Tables
1-4 through 1-9. The tables are organized according to treat-
ment category (biological, physical/chemical, and contain-
ment) and environment (vadose zone and saturated zone).
Each table also is organized according to characterization
parameters, comments, and text reference by sections in the
Handbook. These tables can be used to quickly locate infor-
mation within the Handbook that relates treatment technolo-
gies to specific site characterization parameters.
1.3 Site Reconnaissance
Site reconnaissance activities include gathering informa-
tion on site layout, history, and records of management.
Aboveground natural and cultural features and industrial pro-
cesses are important aspects of a site that may affect subsur-
face processes and the application of subsurface remediation
technologies. Table 1-10 lists important site conditions that
can be used as part of site characterization for subsurface
remediation. Identification of these features and processes
provides critical information concerning potential vadose-
zone and ground-water quality as well as limitations for the
application of subsurface remediation technologies. A site
visit may reveal the industrial processes or waste sources that
contribute to contamination at a site. Observations of topogra-
phy, buildings, parking lots, and waste facilities provide valu-
able information on accessibility for sampling, culturally
induced flow of gases (e.g., beneath buildings), and limita-
tions or constraints to the application of subsurface treatment
technologies (e.g., site size constraints or natural boundaries).
Information on past waste management practices that
documents conditions under which hazardous waste has been
managed is important to site characterization. Table 1-11 lists
important waste management data and records that can be
used in planning a site characterization effort. These records
may include available history of waste disposal and waste
composition. This information may be used in conjunction
with subsurface core and pore-liquid characterization data to
determine areas of contamination and areas of nonhomogeneity,
to evaluate the areal and depth extent of contamination, and to
modify a site characterization plan.
Figure 1-3 presents a flow chart demonstrating an itera-
tive approach for data collection from site characterization
activities for subsurface remediation evaluation and selection,
as well as field optimization of remediation technologies. This
approaches combines site reconnaissance information with
site characterization and sampling, utilizing the methodology
presented in Figure 1-1, for selecting, evaluating, and apply-
ing the subsurface remediation techniques addressed in Tables
1-4 through 1-9.
Table 1-2. Subsurface-Baaed Waste Characterization
Parameter Text Reference (Section)
Chemical class'
Chemical
properties?
Chemical
reactivity
Sorptiorf
Degradation*
Volatilization?
Interphase phase
potential
(7.4.2) (8.1.2) (8.3) (9.1.4) (9.3.1) (9.3.3)
(9.3.4) (12) (14.1) (14.2)
(8.1) (8.3) (10.2) (10.4) (14.1)
(2.2) (7.4.2) (8.1) (8.3) (8.4) (12.1.2)
(12.1.3) (14.1)
(2.2) (8.1.2) (10.3) (12.1.4) (14.1) (15.2.3
(15.2.4)
(2.2) (8.1) (8.3) (13) (15.2.2) (14.1) (16.3;
(5.2.5) (5.4) (9.1.6) (9.2) (i 1) (14.1) (15.2
(5.2.5) (14.1) (14.4)
"Organic (acid, base, polar neutral, nonpolar neutral), and inorganic.
'Molecular weight, melting point, specific gravity, structure, volubility,
ionization, cosolvation.
°Oxidation, reduction, hydrolysis, precipitation, dissolution,
polymerization.
'Adsorption, desorption, ion exchange.
'Biotic, abiotic.
'Henry's Law partitioning, soil gas analysis, vacuum extraction.
"Includes gas, inorganic mineral solid, organic matter solid, water,
and nonaqueous phases.
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Organic (9) (10) (13) (14) (15) (16)
(4) (5) (6) (15) Water
Fluid Phase
Gas
(5) (11) (15)
Solid Phase
Inorganic
(3) (8) (15)
NAPL (10) (16)
Figure 1-2. Problem assessment for site characterization baaed on mass balance approach (Chapters 2, 12, and 14).
Assessment of distribution, reaction, and
migration potential by use of site
reconnaissance' information, site
characterization and sampling, and mass
balance analysis utilizing assessment/
predictive mathematical modeling
I
, I ,
YES I
Information sufficient to demonstrate '
1 distribution, reaction, and migration, and I
treatment potential?
NO\
YES
Additional assessment of use of
laboratory/field studies of distribution,
reaction, and migration and studies of
effects of design and management
parameters on treatment performance
NO
Field verification studies
to monitor treatment
effectiveness
. I Information sufficient to '
I 1 demonstrate treatment \ 4
I optimization? i
' Site reconnaissance activities at a site include gathering information such as
site layout, history, and records of management.
2 Treatment = biological, physical/chemical, or containment (Table 1-3).
Figure 1-3. Flow chart for evaluation of site characterization for subsurface remediation.
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Table 1-3.
Summary of Tables of Characterization
Parameters for Subsurface Remediation
Technologies
Treatment Technology Category
Subsurface
Environment
Physical/
Biological Chemical Containment
[Text Reference (Table Number)]
Vadose zone
Saturated zone
1-4
1-5
1-6
1-7
1-8
1-9
Table 1-4. Characterization for Biological Treatment of Soil in the Vadose Zone"
Parameter
Comments
Text Reference (Section)
Physical
moisture
temperature
permeability
pH
oxygen availability
interphase transfer
potential
Chemical
individual
chemicals
redox potential
C:N:P ratio/
nutrient
Biological
kinetics/activity
enumeration
toxicity
metabolism
affects microbial activity/kinetics
affects microbial activity/kinetics
affects nutrient supply and gas exchange
affects chemical form and microbial activity
affects aerobic/anaerobic metabolism,
activity/kinetics
used in mass balance to determine abiotic removal (56.2.5) (14.1) (14.4)
(4.2.2) (5.2.3) (5.2.4) (5.2.5) (9.2.2) (13.2.2) (14.2.1)
(14.3.2) (15.2.1) (15.2.2) (15.3)
(13.2.2) (14.2.1) (14.3.2) (15.2.2)
(5.2.4) (13.3.1) (14)
(13.2.2) (8.1.2) (12.2.2) (14.2.1) (14.3.2)
(5.2.5) (11) (12.1.3) (13.2) (13.3) (14.2.1) (14.3.2) (15.2.2)
affects rate and extent of degradation
often controlled by microorganisms and
related to aerobic/anaerobic pathway
affects microbial growth
affects rate of degradation
related to population or mass of microorganisms
affects rate and extent of degradation
influences production of (toxic) intermediates
and indicates mechanism(s) of biodegradation
treatability studies can indicate potential for degradation and
important factors controlling rate and extent
adaptation ability of system to acclimate, indicated by
increase in rate and extent of degradation with
incubation time and with repeated applications
of contaminated material
(8.1.2) (13.2.1) (13.3.2) (14.2.2) (15.2.2)
(8.1.2) (13.2.2)
(13.2.2) (13.3.1)
(13.2.4) (13.3.1) (15.2.2)
(13.2.1) (13.3.1)
(13.3.1) (15.2.2)
(13.2.3) (15.2.2)
(14.3.2)
(13.2.3) (15.2.2)
'Approaches and specific techniques for treatment are addressed in Chapters 14 and 15 and listed in Tables 15-3 and 15.4.
-------
Table 1-5. Characterization for Biological Treatment of Aquifer Material in the Saturated Zone*
Parameter
Comments
Text Reference (Section)
Physical
temperature
permeability
geology
geochemistry
pH
oxygen availability
interphase transfer
potential
Chemical
individual chemicals
redox potential
C:N:P ratio/nutrient
Biological
kinetics/activity
enumeration
toxicity
metabolism
treatabiiity studies
adaptation
affects microbial activity/kinetics
affects nutrient suppiy and gas exchange
influences heterogeneity, day lenses
influences waste distribution
influences microbial activity
affects chemical form (mobiiity) and
microbial activity
affects aerobic/anaerobic metabolism, and
activity kinetics
used in mass balance to determine abiotic
removal
affects rate and extent of degradation
often controlled by microorganisms and
related to aerobic/anaerobic pathway
affects microbial growth
affects rate of degradation
related to population or mass of microorganisms
affects rate and extent of degradation
influences production of (toxic) intermediates
and indicates mechanism(s) of biodegradation
can indicate potential for degradation and
important factors controlling rate and extent
ability of system to acclimate, indicated by
increase in rate and extent of degradation with
incubation time and with repeated exposure
(13.2.2) (14.2.1) (14.3.2) (15.2.2)
(4.2.3) (13.3.1) (14)
(3.1) (3.2) (3.3) (3.4)
(8.1.2) (8.3)
(13.2.2) (8.1.2) (12.2.2) (14.2.1) (14.3.2)
(11) (12.1.3) (13.2) (13.3) (14.2.1) (14.3.2) (15.2.2)
(5.2.5) (14.1) (14.4)
(8.1.2) (8.3) (13.2.1) (13.3.2) (14.2.2) (15.2.2)
(8.1.2) (13.2.2)
(13.2.2) (13.3.1)
(8.1.2) (13.2.4) (13.3.1) (15.2.2)
(13.2.1) (13.3.1)
(13.3.1) (15.2.2)
(9.3.1) (13.2.3) (15.2.2)
(14.3.2)
(13.2.3) (15.2.2)
* Approaches and specific techniques for treatment are addressed in Chapters 14 and 16.
-------
Table 1-6.
Technique
Extraction
Characterization Parameters
Parameter
physical
particle size distribution
conductivity/permeability
organic matter
moisture content
heterogeneity/layering
depth
soil gas
interphase transfer potential
for Physical/Chemical Treatment of the Vadose Zone*
Comments
affects volume reduction, sorption, extraction difficulty
affects flow velocity (time) for extraction
affects distribution and sorption of chemicals
affects conductivity of air through soil for vacuum
extraction
affects relative rates of extraction for different layers
along with area, determines volume of contaminated
material and engineering strategies for extraction
used along with soil core analysis to monitor extent
and rate of vacuum extraction
used in mass balance to determine treatment
effectiveness
Text Reference (Section)
(15) (15.2.4)
(3.2)
(5.2.4) (15.2.4)
(3.2)
(4.2.2) (5.1) (5.2.3)
(3.1) (3.3)
(3.5.1) (5.1)
(5.2.5)
(5.2.5) (14.1) (14.4)
Oxidation/
Reduction
Solidification/
Stabilization
chemical
individual chemicals
pH changes
chemicai characteristics'
cation exchange capacity
organic and metal content
redox potential
individual chemicals
redox potential
individual chemicals/sites
porosity/permeability
examples of chemicals that have been treated
may indicate precipitation or dissolution that affects
ease of extraction (permeability)
aids in selection of extraction fluid
determine cation sorption potential, related to clay
content
determine target and/or interfering constituents,
pretreatment needs, extraction fluid
indicates mobile and immobile forms of chemicals
examples of chemicals that have been treated
status of the system before treatment
examples of chemcals/sites that have been treated
affects delivery and mixing of chemicals
(5.4) (15)
(8.1.2)
(5.1.5) (5.4) (7.4.2) (8.2)
(10.3.1) (10.4) (11.1) (11.2.3)
(12. 1)(12.3)
'Approaches and specific techniques for treatment are addressed in Chapters 14 and 15 and are listed in Tables 15-1, 15-2, 15-3, and 15-4.
"Extraction techniques include aqueous, solvent, critical fluid, vacuum (air/steam), and low temperature thermal stripping.
0 Chemical characteristics include vapor pressure, solubility, Henry's Law constant, partition coefficient, boiling point, and specific gravity.
-------
Table 1-7.
Tech-
nique
Characterization Parameters for Physical/Chemical Treatment of the Saturated Zone
Parameter Comments Text Reference (Section)
Product
Removal
Pump-and-
Treat
physical
particle size distribution
in vadose zone
particle size distribution
in saturated zone
flow characterization
geology
organic matter
interphase transfer
potential
chemical
individual chemicals,
contaminants
red ox
soil gas analysis
properties *
physical
particle size distribution
in saturated zone
flow characterization
geology
organic matter
interphase transfer
potential
chemical
individual chemicals,
contaminants
red ox
soil gas analysis
propertied
organic-inorganic
interactions
affects amount of contaminant stored
in capillary fringe for LNAPL
affects permeability and product
retention
affects direction, location, and
extent of LNAPL
influences distribution of DNAPL
and LNAPL
affects distribution and sorption
assists in determining phase (s) where
more than one phase is involved
examples of contaminants that have
been treated
temporal and spatial variation may
influence permeability
assist in locating contamination (area)
assist in locating contamination (depth)
affects pumping (recovery) rate of
water and contaminant
affects direction, location, and
extent of contamination
influences distribution of contaminants
affects distribution and sorption
assists in determining phase(s) where
contaminant is found
examples of contaminants that have
been treated
temporal and spatial variation may
influence permeability
assist in locating contamination (area)
assist in locating contamination (depth)
affects design of systems
(3.2)(10.2)
(2.2)(2.4)(10.2)
(3.2)(10.3.2)
(8.1.2)(8.3.3)(8.4)(9.1.3)(13.2.2)
(5.2.5) (9. 2.1)
(2.2)(2.3)(2.4)(4.1)(4.2.2)
(3)(6.4)
(3.2)(10.3.2.)
(3.5.2)(4.4)(4.5)
(8.1.2)(8.3.3)(8.4)(9.1.3)(13.2.2)
(5.2.5) (9.2.1)
(10.2)
(12.3)
Approaches and specific techniques for treatment are addressed in Chapters 14 and 16, Section 16.1.
' Properties include molecular weight, specific gravity, volubility, melting point, structure, ionization, and cosolvation.
-------
Table 1-8. Characterization Parameters for Immobilization''/Containment"Techniques in the Vadose Zone'
Technique
Parameter
Comments
Text Reference (Section)
Immobilization
physical
particle size distribution
moisture content
permeability
organic matter
depth
lithology
interphase transfer
potential
affects sorption, ion exchange
affects efficiency, energy requirements,
and sorption
affects delivery of chemicals
affects distribution and sorption
along with area, determines volume of
contaminated material and engineering
strategies
affects extent of sorption and ion
exchange
used in mass balance to evaluate
solution to solidphase transfer for
immobilization
(3.2)
(5. 1)(5.2.3)(14. 1)(15)(15.2.3)
(5.2.4)(15.2.3)
(3.2)
Containment
chemical
individual chemicals
contaminants
redox potential
pH
cation exchange capacity
properties
physical
stratigraphy
interphase transfer
potential
containment requirements
examples of contaminants that have
been treated
affects chemical speciation and thus
immobilization
affects chemical speciation and thus
immobilization
affects ion exchange
affects affinity of chemicals for
surfaces and for precipitation
identify path ways and extent of
chemical migration
used in mass balanceto evaluate
success of containment
evaluate containment of gas, liquid,
and solid phases
(5.5)
(8.1.2)
(8.1.2)
(8.1.2)
(8. 1)(8.3)(10.3. 1)(10.3.2)(10.4)
(3.1)
(5.2.5)(14.1)(14.2.2)(14.2.3)(14.4)
(14.2.3)
immobilization techniques include sorption, ion exchange, precipitation, stabilization/solidification, and vitrification.
' Containment techniques include physical stuctures.
! Approaches and specific techniques for treatment are addressed in Chapters 14 and 15, Section 15.2.3.
1 Properties include molecular weight, melting point, specific gravity, structure, ionization, solubility, and cosolvation.
-------
Table 1-9. Characterization Parameters for
Technique Parameter
Containment Techniques" in the Saturated Zone
Comments
Text Reference (Section)
Hydraulic
Physical
Structuresf
physical
flow system characterization
permeability
geology
advection
interphase transfer
potential
fracture flow
physical gradients
chemical
contaminants present
individual chemicals/
contaminants
environmental parameters'
chemical gradients
propeties"
reactions'
physical
flow system characterization
geology
fracture flow
determine area and depth for
containment
affects rate of movement and rate of
pumping
assists with flow system
characterization
generally primary transport (escape)
path
used in mass balance to assess and
evaluate containment
may exercise control on ground-water
flow
affects geochemistry, which may affect
permeability
identify chemicals of concern that
might escape
examples of contaminants that have
been contained
may change with pumping and affect
recovery and permeability
may affect geochemistry if reinfected
and affect permeability
affects affinity of chemicals for
surfaces and for precipitation, as well
as interphase transfer
may affect treatment/permeability while
pumping
determine area and depth for
containment
assists with flow system
characterization
may exercise control on ground-
water flow
(2.2)
(4.4)
(3)
(4.1)
(5.2.5)(14.
(6.1)
(8.3.2)
(2.3)
(3.5.2)(4.4)(4.5)
(8.1.2)
(8.3.3)
(2.2)
(3)
(6.1)
chemical
contaminants present
individual chemicals/
contaminants
identify chemicals of concern that
might escape
examples of contaminants that have
been contained
(2.3)
(3.5.2)
'Containment techniques may be temporary and used as part of a treatment train that includes product removal, pump-and-treat, pumping
and reinjection, and bioremediation.
'Approaches and specific techniques are addressed in Chapters 4 (section 4.4), 14 and 16 (Section 16.2).
0Environmental parameters include pH, alkalinity, redox potential, salinity, temperature, and pressure.
"Properties include molecular weight, melting point, specific gravity, structure, solubility, ionization, and cosolvation.
'Reactions include hydrolysis, substitution, elimination, oxidation-reduction, and biodegradation.
"Physical structures often are used in conjunction with hydraulic containment and withdrawal (e.g., clay cap to reduce recharge combined with
extraction wells to remove chemical) (refer to Section 3.5.2).
10
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Table 1-10. Aboveground Features for Site Characterization
Item Specific Information
Site location
Climatological data
Topographic map, including contours, map scale and date, floodplain areas, surface waters, springs and
intermittent streams, and site legal boundaries.
Site map, including injection and withdrawal wells on site and off site; buildings and recreation areas, access
and internal roads; storm, sanitary, and process sewerage systems; loading and unloading areas; and fire
control facilities.
Location of past ano/or present operation units and equipment cleaning areas, ground-water monitoring wells,
delineation of waste management units, and site modifications.
Surrounding area land use patterns.
Vegetation (trees, shrubs, grasses).
Precipitation/evaporation/humidity.
Site water budget.
Temperature (averages and extremes)
Wind rose.
Predicted storm events (e.g., 24-hour, 25-year, average number of days of rain and snow).
Frost action potential
Table 1-11. Waste Management Information for Site Characterization
Category Item
Specific Information
History of waste application
Years in operation and annual
quantity of waste generated
and/or disposed.
Records of measured annual waste quantity (weight/volume)
over life of site. Include hazardous and nonhazardous
managed at same site.
History of waste quality
Placement of waste.
Size of waste unit(s)
Waste analyses.
Unit processes.
Disposal areas.
Records of quantity (weight/volume), and location of each
waste disposal action.
Area and depth.
Periodic analyses of hazardous, wastes.
History of unit processes employed in the generation and
treatment of wastes.
Pits, ponds, lagoons, landfills, storage tanks, wastewater
treatment plant locations (present and historical).
II
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PART I: METHODS FOR SUBSURFACE CHARACTERIZATION
Chapter 2
Site Characterization Overview
James W. Mercer and Charles P. Spalding
2.1 Introduction
Characterization of a hazardous waste site involves gath-
ering and analyzing data to describe the processes controlling
the transport of wastes from the site. It provides the under-
standing to predict future site behavior based on past site
behavior. It can encompass the characterization of the waste
itself as well as that of various transport pathways such as air,
surface water, biota, and ground water. Ground water, the
focus of this discussion, is often the most significant and least
apparent transport pathway.
Site characterization follows the scientific method and is
performed in phases (see Figure 2-1). First, a hypothesis is
made concerning site or system behavior. Based on this
hypothesis, a data collection program is designed, data are
collected, and an analysis or assessment is made. Using the
results of the analysis, the hypothesis is refined and additional
data may be collected. As the knowledge of the site becomes
more detailed, the working hypothesis may take the form of
either a numerical or analytical model. Data collection contin-
ues until the hypothesis is proven sufficiently to form the
basis for decision making.
Because the ultimate goal of site characterization is to
make informed decisions, the first step is to define study
objectives. A possible list of objectives, provided by Cartwright
and Shafer (1987), includes the following: (1) assess the
background or "ambient" water quality (how was the water
before contamination?); (2) establish the impacts of certain
facilities, practices, or natural phenomena on water quality
(what is the extent of contamination?); and (3) predict future
ground-water quality trends under a variety of conditions
(what would be the impact of various remedial actions?).
Whatever the objectives, ground-water site characteriza-
tion has two major components: assessment of the ground-
water flow system and assessment of the contamination in the
ground water. All too often, emphasis is placed on the latter
component, which involves ground-water quality monitoring,
Everett (1980) defines monitoring as a scientifically designed
surveillance system of continuing measurement and observa-
tions. At many waste sites, ground-water quality data are
abundant; however, water-level data used to determine advec-
tive transport are limited. This is unfortunate because water-
level data are equally important, and they are easier and less
expensive to collect than water-quality data.
This chapter provides an overview of Part I of the Hand-
book, which focuses on methods of site characterization. This
chapter covers the following topics (1) flow system charac-
terization, (2) contamination characterization, (3) techniques
for characterization, and (4) analysis of data.
2.2 Flow System Characterization
Flow system characterization begins with an understand-
ing of controlling processes and of the data required to define
those processes (Table 2-1). Ground water is always in motion
from areas of natural and artificial recharge to areas of natural
and artificial discharge. Natural recharge occurs from precipi-
tation and surface water bodies; artificial recharge results
from human-induced actions such as irrigation and well injec-
tion. Ground water discharges naturally to springs and other
surface water bodies, e.g., rivers, lakes, and oceans. Under
natural conditions, ground water moves very slowly, its flow
velocity ranging from a fraction of a foot per year to several
feet per day. In most cases, flow obeys Darcy's law, which
states that the velocity is proportional to both the hydraulic
conductivity of the formation and the hydraulic gradient.. The
term hydraulic conductivity is used to express the water-
conducting capacity of the formation material. The hydraulic
gradient is an expression of the slope of the ground-water
surface.
Shallow aquifers are usually important sources of ground
water. These upper aquifers are also the most susceptible to
contamination. Contaminants may enter an upper aquifer in
oneof the following ways: (1) artificial recharge or leakage
through wells; (2) infiltration from precipitation or irrigation
13
-------
Analyze and test
hypothesis
Figure 2-1. Site characterization phasea (from Bouwer et al.,
1988).
return flow through the vadose zone above the water table; (3)
induced recharge from influent streams and lakes or other
surface water bodies; (4) inflow through aquifer boundaries
and leakage from overlying or underlying formations; and (5)
leakage or seepage from impoundments, landfills, or miscel-
laneous spills.
Water and contaminants carried with it may leave an
aquifer in the following ways: (1) ground-water leakage from
the aquifer into adjacent strata, (2) ground-water withdrawal
by pumping and drainage, (3) seepage into effluent streams
and lakes, (4) spring discharge, and (5) evapotranspiration.
Data required to assess these processes are shown in Table 2-
2. In general, these data requirements include a geometric
description of the site (layering and hydraulic boundaries);
storage and transmissive properties; and source/sink informa-
tion, such as wells. More specific lists of data with ranges of
values are provided in Mercer et al. (1982). For any particular
site, it is rare to have all this information. Data gaps can be
addressed by a field collection program, but to some extent
must be filled based on experience. In addition to physical and
chemical data, other factors listed in Table 2-2 include regula-
tory and legal issues such as water rights and future land use.
The first step in designing a field program is to review
existing data for the site or nearby locations. Sources of
information include the U.S. Geological Survey (USGS) (Mer-
cer and Morgan, 1981); state geologic and water agencies;
local water districts; and city, county, and state health depart-
ments. Other federal agencies that may provide data include
the U.S. Environmental Protection Agency (EPA) (e.g., the
STORET computerized information storage system); U.S.
Bureau of Reclamation; U.S. Army Corps of Engineers; and
U.S. Soil Conservation Service. Additional inforation may
be available from consultants and universities. Several data
sources are discussed below.
The U.S. Department of Agriculture, Soil Conservation
Service, has three soil geographic data bases the Soil Survey
Geographic Data Base (SSURGO), the State Soil Geographic
Data Base (STATSGO), and the National Soil Geographic
Data Base (NATSGO). Components of map units in each
geographic data base are generally phases of soil series. The
Soil Conservation Service also maintains a soil interpretations
record data base, which encompasses more than 25 soil,
physical, and chemical properties for the 15,300-plus soil
series recognized in the United States. Interpretations are
displayed differently for each geographic data base to be
consistent with the level of detail expressed. Particle size
distribution, bulk density, available water capacity, soil reac-
tion, salinity, and organic matter are included for each major
layer of the soil profile. Data on flooding, water table, bed-
rock, and subsidence characteristics of the soil; and interpreta-
tions for erosion potential, septic tank limitations, engineering,
building and recreation development, and cropland, wood-
land, wildlife habitat, and rangelands management also are
included.
The U.S. Department of Interior Geological Survey cre-
ated and maintains a central storage facility for water re-
sources data, known as the National Water Data Storage and
Retrieval System (WATSTORE), at its National Headquar-
ters in Reston, Virginia. Included in this computerized storage
facility are representative ground-water data collected through-
out the United States, This ground-water information resides
in a computer data file, which is maintained by a database
management system (DBMS) called SYSTEM 2000. The
name and acronym given this data base is the Ground-Water
Site-Inventory (GWSI) file. Although several field-collected
parameters of water-quality data (including temperature, con-
ductance, and pH) are stored in the GWSI, the bulk of water-
quality data reside in a nationwide file called Storage and
Retrieval (STORET), a file maintained by EPA. The National
14
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Table 2-1. A Summary of the Processes Associated with Dissolved Solute Transport and Their Impact
Process Definition Impact on Transport
Solute Transport
Advection
Diffusion
Dispersion
Solute Transfer
Radioactive decay
Sorption
Movement of solute as a consequence of ground-water flow.
Solute spreading due to molecular diffusion in response to
concentration gradients.
Fluid mixing due to effects of unresolved heterogeneities in
the permeability distribution.
Irreversible decline in the activity of a radionuclide through a
nuclear reaction.
Most important way of transporting solute away
from source.
An attenuation mechanism of second order in
most flow systems where advection and
dispersion dominate.
An attenuation mechanism that reduces solute
concentration in the plume. However, it
spreads to a greater extent than a plume
moving by advection alone.
Partitioning of a solute between the ground water and mineral
or organic solids in the aquifer.
An important mechanism for attenuation when the
half-life for decay is comparable to or less
than the residence time of the flow system.
Also adds complexity in production of
daughter products.
An important mechanism that reduces the rate at
which the solute is apparently moving. Makes it
more difficult to remove solute at a site.
Dissolution
precipitation
Acid-base
reactions
Complexation
Hydrolysis/
substitution
Redox reactions
(biodegradation)
Biologically Mediated
Mass Transfer
Biological transfor-
mations
The process of adding solutes to or removing them from solution Precipitation is an important attenuation
by reactions dissolving or creating various solids. mechanism that can control the concentration in
solution. Solution concentration is mainly
controlled either at the source or at a reaction
front.
Reactions involving a transfer of protons (I-P).
Combination of cations and anions to form more complexion.
Reaction of a halogenated organic compound with water
or a component ion of water (hydrolysis) or with
another anion (substitution).
Reactions that involve a transfer of electrons and
include elements with more than one oxidation state.
Reactions involving the degradation of organic compounds
and whose rate is controlled by the availability of
nutrients to adapted microorganisms and redox conditions.
Mainly an indirect control on solute transport by
controlling the pH of ground water.
An important mechanism resulting in increased
volubility of metals in ground water, if adsorption
is not enhanced. Major ion complexation will
increase the quantify of a solid dissolved in
solution.
Often hydrolysis/substitution reactions make an
organic compound more susceptible to
biodegradation and more soluble.
An extremely important family of reactions in
retarding solute spread through precipitation of
metals.
Important mechanism for solute reduction, but can
lead to undesirable daughter products.
From NRC, 1990
Water Data Exchange (NAWDEX) Local Assistance Centers
are authorized users of the STORE! file and may retrieve
ground-water quality data for subscribers.
A field program usually follows a data review of hydro-
geologic investigation techniques (U.S. EPA, 1986 and Sisk,
1981). Summaries of procedures for well installation and
aquifer testing are described in Ford et al. (1984) and Aller et
al. (1989). Kruseman and de Ridder (1976), Lawrence Berke-
ley Laboratory (1977, 1978) discuss methods of analysis of
aquifer and slug tests. In general, as the scale of the observa-
tion increases, the range of measured properties, such as
hydraulic conductivity, tends to change because of the hetero-
geneous nature of geologic materials. Particularly, ground-
water flow rates estimated from measurements on cores may
underestimate ground-water flow rates in the area if flow is in
fractures or in other more permeable layers.
Because of seasonal changes in ground water, a minimum
of one year should be devoted to characterization. As the site
complexity increases, this period will increase proportion-
ately. Several factors influence the number of boreholes re-
quired, the most important being heterogeneities in the aquifer
materials. Methods of quantifying ground-water networks are
not widely used but do exist. For example, van Geer (1987)
shows how Kalman filters are used to design ground-water
monitoring networks. Another technique used to evaluate
ground-water networks is kriging (e.g., Olea, 1982); this
technique is discussed further in Section 2.5 and in Chapter 7
(Section 7.3.2).
15
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Table 2-2.
Data Pertinent to the Prediction of Ground-Water Flow
I. Physical Framework
1. Hydrogeologic map showing areal extent and boundaries of aquifer.
2. Topographic map showing surface-water bodies.
3. Water-table, bedrock-configuration, and saturated-thickness maps.
4. Hydraulic conductivity map showing aquifer and boundaries.
5. Hydraulic conductivity and specific storage map of confining bed.
6. Map showing variation in storage coefficient of aquifer.
7. Relation of stream and aquifer (hydraulic connection).
II. Stresses on System
1. Type and extent of recharge areas (irrigated areas, recharge basins, recharge wells, impoundments, spills, tank leaks, etc.).
2. Surface water diversions.
3. Ground-water pumpage (distributed in time and space).
4. Stream flow (distributed in time and space).
5. Precipitation and evapotranspiration.
III. Observable Responses
1. Water levels as a function of time and position.
IV. Other Factors
1. Economic information about water supply.
2. Legal and administrative rules.
3. Environmental factors.
4. Planned changes in water and land use.
After Moore, 1979
2.3 Contamination Characterization
As with flow system characterization, contamination char-
acterization begins with understanding the processes control-
ling transport and degradation (Table 2-1) and the data required
to define those processes. These processes determine mini-
mum data requirements needed to characterize a site.
Nonreactive (conservative) dissolved contaminants in satu-
rated porous media are controlled by the following factors:
l.Advection: This mechanism causes contaminants to
be transferred by the bulk motion of the ground
water. The term convection is sometimes used in
place of advection.
2. Mechanical (or kinematic) dispersion: This process
involves mechanical mixing caused by three mecha-
nisms. The first mechanism occurs in individual pore
channels because molecules travel at different ve-
locities depending on whether they are near the edge
or in the center of the channel. The second mecha-
nism is triggered by differences in surface area and
roughness relative to the volume of water in indi-
vidual pore channels, causing different bulk fluid
velocities in different pore channels. The third mecha-
nism is related to the tortuosity, branching, and
interfingering of pore channels, causing the stream-
lines to fluctuate with respect to the average flow
direction. Mechanical dispersion occurs in the direc-
tion of the average flow velocity and in the plane
orthogonal to the average flow direction, These ef-
fects are called longitudinal dispersion and trans-
verse dispersion, respectively, Longitudinal dispersion
is due to variations of the velocity component along
the average flow direction, whereas transverse dis-
persion is due to variations of the velocity compo-
nents in the normal plane.
3. Molecular diffusion: Fickian diffusion causes the
contaminant molecules or ions to move from high
concentrations to lower concentrations. Movement
also is caused by the random kinetic motion of the
ions or molecules (Brownian diffusion).
The combined effect of mechanical dispersion and mo-
lecular diffusion is known as hydrodynamic dispersion. Dis-
persion causes the zone of contaminated ground water to
occupy a greater volume than if the contaminant distribution
were influenced only by advection. If a slug of contaminant
enters the ground-water system, advection causes the slug to
move in the direction of ground-water flow. Hydrodynamic
dispersion causes the volume of the contaminated zone to
increase and the maximum concentration in the slug to de-
crease. Transverse dispersion may expand a contaminant plume
10 to 20 percent beyond the width defined by convective
transport (Lehr, 1988). Macroscopic variations in hydraulic
conductivity and porosity are probably more significant fac-
tors affecting solute transport than hydrodynamic 1 dispersion
changes (Wheatcraft, 1989).
Additional processes affect transport for reactive con-
taminants. In addition to advection and hydrodynamic disper-
sion, the migration of reactive contaminants is further controlled
by adsorption, desorption, chemical reactions, and biological
transformation.
16
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I. Adsorption or desorption: These processes involve
mass transfer of contaminants. Adsorption is the
transfer of contaminants from the ground water to
the soil. Resorption is transfer of contaminants from
the soil to the ground water.
2. Chemical reactions: These processes involve mass
transfer of contaminants caused by various chemical
reactions (e.g., precipitation and dissolution, oxida-
tion and reduction). For some contaminants, degra-
dation is also an important process that may need to
be characterized.
3. Biological Transformation: These processes may re-
move contaminants from the system by biological
degradation, or transform contaminants to other toxic
compounds that are subject to mass transfer by the
other processes discussed above.
The processes of adsorption-desorption, chemical reac-
tions, and biological transformation play important roles in
controlling the migration rate as well as concentration distri-
butions. These processes tend to retard the rate of contaminant
migration and act as mechanisms to reduce concentrations.
Because of their effects, the plume of a reactive contaminant
expands and the concentration changes more slowly than
those of an equivalent nonreactive contaminant (see Figure 2-
2). As discussed in subsequent chapters, however, resorption
can require longer time periods to reach concentration cleanup
standards.
Table 2-3 shows data requirements for contamination
characterization, in addition to the requirements shown in
Table 2-2. For example, to characterize advective transport,
the flow system must first be understood. More specific lists
of data with ranges of values are provided in Mercer et al.
(1982). These data requirements provide a broad view of the
factors affecting contaminant transport from a site.
2.4 Techniques for Characterization
For site characterization, it is important to understand the
transport mechanisms and ground-water flow system at a site.
Once these mechanisms and systems are understood, ground-
water monitoring data can be interpreted to obtain information
far more useful than simple information on contaminant levels
at specific points and times. The procedures used to obtain
water-quality data are of critical importance. Procedures for
drilling monitoring wells, taking samples, and having samples
analyzed by a laboratory are discussed in this section.
Table 2-4 shows actions that were typically taken at
hazardous waste sites in the early 1980s. Two data gaps are
the vertical distribution of hydraulic head, as measured by
water levels in adjacent wells cased to different depths, and
hydraulic conductivity values. Therefore, most guidance docu-
ments now recommend the actions shown in Table 2-5. At
sites where conditions warrant (e.g., fractured media), addi-
tional actions may be necessary to fully characterize the site
(see Table 2-6).
A variety of common well drilling methods maybe used
to install monitoring wells at hazardous waste sites. These
methods include solid stem continuous flight and hollow stem
continuous flight augering, cable tool drilling, mud and air
rotary drilling, jetting, and driving well points. Detailed dis-
cussions of the principles of operation of each of these meth-
ods are available from numerous sources including Scalf et al.
(1981), Driscoll (1986), and Campbell and Lehr (1973). A
summary of the advantages and disadvantages of various
drilling methods relative to monitoring well construction is
provided in Scalf et al. (1981) and Larson (1981), as well as in
Chapter 4 of this Handbook (Section 4.2.1).
A variety of materials are available for use in casing,
screening, and other structural and sampling components of
monitoring wells. The most commonly used are mild steel,
stainless steel, polyvinyl chloride (PVC), polypropylene, poly-
ethylene, and Teflon®. Barcelona et al. (1983) summarizes the
characteristics of several of these materials. These materials
have substantially different properties relative to strength,
corrosion resistance, interference with specific contaminant
measurements, expense, and availability. Consequently, they
must be selected carefully and demonstrated to be the most
appropriate for the particular monitoring program. Consider-
ations should include all pertinent, site-specific factors such
as well installation method, depth, geochemical environment,
and probable contaminants to be monitored. Well casing
materials are discussed further in Section 4.2.1 (see especially
Table 4-3) and Section 9.3.4.
Construction details for individual wells should be docu-
mented and verifiable through the use of drilling logs. The
drilling log should contain information about the texture,
color, size, and hardness of the geologic materials encoun-
tered during the drilling (Barcelona et al., 1985). Any use of
drilling fluids, grouts, and seals also should be noted in the
record of well construction. Well casing materials should be
documented because the type of well casing may have an
effect on the quality of the water samples (Barcelona et al.,
1983). The same considerations that apply to well casing
materials for newly constructed monitoring wells apply to
evaluating the suitability of existing wells for ground-water
quality monitoring.
Guidance documents on ground-water monitoring em-
phasize the need for depth-discrete data to determine the
three-dimensional flow field and chemical distribution
(Barcelona et al., 1983; Barcelona et al., 1985; and U.S. EPA,
1986). Shorter well screens and more nested wells are recom-
mended where immiscible liquids (liquids that tend to float
above water or sink to the bottom), heterogeneous conditions,
or a thick flow zone are present (U.S. EPA, 1986). Barcelona
et al. (1983) recommend installing nested wells with short
well screens (less than 5 ft long) where the potential flow zone
is more than 10 ft thick.
Once the wells are designed and drilled, accepted practice
is to remove fluid from the formation, with subsequent labora-
tory analysis of the sample (Morrison, 1983; de Vera, 1980;
USATHAMA, 1982 Guswa et al., 1984; and Everett et al.,
1984). This approach results in a set of point data that repre-
sent (depending on the type of well construction, the sampling
17
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A Advection
D Dispersion
S Sorption
B Biotransformation
A -i- D + S + B
•A + D
Distance from Continuous Contaminant Source
I
A+D+S+B
.A + D
Distance from Slug-Release Contaminant Source
Figure 2-2. The influence of natural processes on levels of
contaminants downgradient from continuous and
slug-release sources (from Keety et al., 1986).
mechanism, laboratory procedure, and hydrodynamics of the
ground-water system), particular aspects of the in situ water
quality at a specific time. Much work (Gibb et al., 1981;
Gillham et al., 1983; Keith et al., 1983; Nacht, 1983; Barcelona
et al., 1984; Olea, 1984; Barcelona et al., 1985) has focused
on improving this process (i.e., providing greater quality
control and quality assurance). Chapter 9 discusses sampling
of subsurface solids and ground water in more detail.
2.5 Analysis of Data
Although this section emphasizes network design and
sampling considerations, no section on data analysis would be
complete without a discussion of database management sys-
tems (DBMS) and geographic information systems (GIS). At
hazardous waste sites, large amounts of data are generated. To
take full advantage of these data in the interpretation stage,
they should be in electronic/magnetic format for use in a
Table 2-3. Data Pertinent to Prediction of the Pollutants in
Ground Water (in addition to those in Table 2-2)
I. Physical Framework
1. Estimates of the parameters that comprise hydrodynamic
dispersion.
2. Effective porosity distribution.
3. information on natural (background) concentration
distribution (water quality) in the aquifer.
4. Estimates of fluid density variations and relationship of
density to concentration (most important where
contaminant is salt water or results in significantly higher
concentration of total dissolved solids compared to the
natural aquifer or where there are significant temperature
differences between the contaminant plume and the
natural aquifer).
II. Stresses on System
1. Sources and strengths of pollutants,
III. Chemical/Biological Framework
1. Mineralogy media matrix.
2. Organic content of media matrix.
3. Ground-water temperature.
4. Solute properties.
5. Major ion chemistry.
6. Minor ion chemistry.
7. Eh-pH environment.
IV. Observable Responses
1. Areal and tamporal distribution of water quality in the
aquifer.
2. Stream flow quality (distribution in time and space)
DBMS and/or GIS. Both systems can be used to manipulate,
correlate, and display data, and this method of organizing
large amounts of data facilitates the interpretation process.
The assessment of ground-water quality on any scale
involves the estimation of chemical variables distributed in
three-dimensional space. A key consideration in establishing
an effective and efficient ground-water quality monitoring
program is the spatial distribution of sampling locations. Care
must be taken in designing monitoring well networks to avoid
biasing any inferences made from the resulting data.
As pointed out, knowledge of the hydrodynamics of the
ground-water system(s) being monitored is also of critical
importance for the design of monitoring networks. For certain
ground-water monitoring program objectives, an optimum
monitoring network for a relatively homogeneous porous
flow environment is different from that for a discretely frac-
tured hydrogeologic medium. For other monitoring objec-
tives, however, the fundamental differences between flow
regimes may have very little impact on the design of an
optimum sampling network.
Proper ground-water sampling and analysis are equally
important for assuring effective ground-water monitoring. A
quality assurance program composed of well-conceived and
effectively implemented quality control procedures should be
followed (Cartwright and Shafer, 1987). Strict adherence to
18
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Table 2-4. Acthons Typically Taken
1. Install shallow monitoring wells.
2. Sample ground water numerous times for a range of pollutants
such as those constituents contained in the RCRA Appendix
IX ground-water monitoring list.
3. Define geology primarily by drillers' logs and drill cuttings.
4. Evaluate local hytrology with water level contour maps of
shallow wells.
5. Possibly obtain soil and core samples for chemical analyses.
Benefits
1. Screening of the site problems is rapid.
2. Costs of investigation are moderate to low.
3. Field and laboratory techniques used are standard.
4. Data analysis/Interpretation is straightforward.
5. Tentative identification of remedial alternatives is possible.
Shortcomings
1. True extent of site problems may be misunderstood.
2. Selected remedial alternatives may not be appropriate.
3. Optimization of final remediation design may not be possible.
4. Cleanup costs remain unpredictable, tend to excessive levels.
5. Verification of compliance is uncertain and difficult.
Modified from Keely et al., 7986
Table 2-5.
Recommended Actions
1. Install depth-specific clusters of monitoring wells.
2. Initially sample for a range of pollutants, but subsequently,
become more selective.
3. Define geology by extensive coring/sediment samplings.
4. Evaluate local hydrology wth well clusters and geohydraulic
tests.
5. Perform limited tests on sediment samples (grain size, clay
content, etc.).
6. Conduct surface geophysical surveys (resistivity, EM, ground-
penetrating radar).
Benefits
1. Conceptual understanding of site problems is more complete.
2. Prospects are improved for optimization of remedial actions.
3. Predictability of remediation effectiveness is increased.
4. Cleanup costs are lowered, estimates are more reliable.
5. Verification of compliance is more soundly based.
Shortcomings
1. Characterization costs are somewhat higher.
2. Detailed understanding of site problems is still difficult.
3. Full optimization of remediation is still not likely.
4. Field tests may create secondary problems (disposal of pumped
waters).
5. Demand for specialists is increased, shortage is a key limiting
factor.
Modified from Keely et al., 7986
quality assurance programs minimizes both systematic and
random errors, and maximizes the likelihood of collecting
ground-water samples in a manner that ensures the reliability
of analytical determinations. As with monitoring network
design, a detailed understanding of the overall objectives of
the monitoring program is a key factor in determining sam-
pling and analysis requirements. See Chapter 7 for further
Table 2-6. Additional Actions Where Conditions Warrant Them
I. Assume Table 2.5 as starting point.
2. Conduct soil vapor surveys for volatiles and fuels.
3. Conduct tracer tests and borehole geophysical surveys (neutron
and gamma).
4. Conduct karst stream tracing and recharge studies, if
appropriate to the setting.
5. Conduct bedrock fracture orientation and interconnectivity
studies, if appropriate.
6. Determine the percent organic carbon and cation exchange
capacity of solids.
7. Measure redox potential, pH, and dissolved oxygen levels of
subsurface.
8. Evaluate sorption-desorption behavior by laboratory column and
batch studies.
9. Assess the potential for biotransformation of specific
compounds.
Benefits
1. Thorough conceptual understandings of site problems are
obtained.
2. Full optimization of the remediation is possible.
3. Predictability of the effectiveness of remediation is maximized.
4. Cleanup costs maybe lowered significantly, estimates are
reliable.
5. Verification of compliance is assured.
Shortcomings
1. Characterization costs may be much higher.
2. Few previous applications of advanced theories and methods
have been completed.
3. Field and laboratory techniques are specialized and are not
easily mastered.
4. Availability of specialized equipment is low.
5. Need for specialists is greatly increased (it may be the key
limitation overall).
Keely etai, 7986
discussion of sources of error in sampling and considerations
in the development of quality assurance programs.
The results of laboratory analyses are only as reliable as
the samples, field standards, and blanks received (Cartwright
and Shafer, 1987). Therefore, to assure that representative
samples are provided to the laboratory, careful thought and
practice must be part of any sampling program. A representa-
tive sample accurately reflects in situ conditions in proximity
to the sample point at the time the sample was collected.
Maintaining representative samples requires consideration of
well purging, sample collection, and sample preservation.
Barcelona et al. (1985) have prepared an extensive guide to
the practical aspects of ground-water sampling. (See also
Chapters 7, 8, and 9 of this Handbook.)
Parameter selection is an important aspect of the design
of a sampling program. The types of hydrochemical measure-
ments to be made affect the choice of sampling equipment and
the sampling methodology. Barcelona et al. (1985) state that it
is often wise to obtain slightly more chemical and hydrologic
data than immediately required in order to aid subsequent
interpretation. Sections 9.2.1 and 9.3.1 discuss further selec-
tion of analytes for the vadose and saturated zones.
19
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The frequency of sample collection is important in the
design of an optimum ground-water quality monitoring pro-
gram (Cartwright and Shafer, 1987). Sampling frequency
affects the cost of the monitoring program and the appropri-
ateness of any inference(s) made from the resulting data.
Sample collection and analysis should not occur so often as to
result in redundant information that would increase costs with
no marginal gain in useful information. Conversely, sample
collection should not be so infrequent as to detract from the
ability to accurately forecast trends in ground-water quality.
Ground-water sampling frequency should be based on the
objectives of the monitoring program and the hydrodynamics
of the ground-water system being monitored. As discussed,
since ground-water movement is relatively slow, there is little
need to sample every few meters of the flow path. Sampling
frequency is discussed further in Section 9.1.4.
During the past decade, the use of geostatistical prin-
ciples (i.e., structural analysis, kriging, and conditional simu-
lation) to interpret ground-water data has increased.
Geostatistical techniques are used to evaluate the spatial vari-
ability of ground-water flow parameters, particularly hydrau-
lic head and transmissiviry. However, less work has been
conducted on the application of geostatistics to interpret
hydrochemical data and ground-water quality monitoring net-
work design. Samper and Neuman (1985), who performed a
geostatistical analysis of selected chemical variables, showed
that geostatistical approaches may be valid to evaluate ground-
water chemical data, particularly on a regional scale (Cartwnght
and Shafer, 1987).
The principles of geostatistics may be appropriate for
interpolation of point data to estimate the spatial distribution
of certain aspects of ground-water quality (Englund and Sparks,
1988). Kriging measures the error of estimation, which can be
mapped and used to select locations for additional sampling
points. These error maps show where the interpolated values
deviated from the expected statistical structure, thus indicat-
ing the best locations to place additional wells (Virdee and
Kottegoda, 1984). However, this information can only serve
as a guide because of other constraints on well location such
as environmental concerns, political issues, and economic
limitations (see Table 2-2). Nevertheless, a near-optimal moni-
toring network can be developed for a predetermined level of
reliability.
The use of geostatistics to design monitoring networks
and interpolate data has limitations. Using kriging for ground-
water investigations often may have a limited effectiveness
because of lack of sufficient data to perform the structural
analysis. Hughes and Lettenmaier (1981) suggest that a mini-
mum sample size of 50 is required before kriging is superior
to more traditional interpolation schemes (e.g., the least squares
method). Even with sufficient data and suitable statistical
support, structural analysis is highly subjective. Further, the
theoretical basis for the application of geostatistics is the
concept of a regionalized variable, which is defined as a
spatially correlated random variable. To date, there have been
no definitive studies of the validity of assuming that hydro-
chemical properties of ground water behave as regionalized
phenomena (Cartwright and Shafer, 1987). For a further dis-
cussion of geostatistical methods, see Section 7.3.2.
2.6 References
Aller, L, T.W. Bennett G. Hackett, Rebecca J. Petty, J.H.
Lehr, H. Sedoris, D.M. Nielsen. 1989. Handbook of
Suggested Practices for the Design and Installation of
Ground-Water Monitoring Wells. EPA/600/4-89/034
(NTIS PB90-159807). Also published in NWWA/EPA
series, National Water Well Association, Dublin, OH.
Barcelona, M.J., J.P. Gibb, and R.A. Miller. 1983. A Guide to
the Selection of Materials for Monitoring Well Construc-
tion and Ground Water Sampling. ISWS Contract Report
327. Illinois State Water Survey, Champaign, IL.
Barcelona, ML, J.A. Helfrich, E.E. Garske, and J.P. Gibb.
1984. A Laboratory Evaluation of Ground Water Sam-
pling Mechanisms. Ground Water Monitoring Review
4(2):32-41.
Barcelona, ML, J.P. Gibb, J.A. Helfrich, and E.E. Garske.
1985. Practical Guide for Ground-Water Sampling. EPA
600/2-85/104 (NTIS PB86-137304). Also published as
ISWS Contract Report 374, Illinois State Water Survey,
Champaign, IL.
Bouwer, E., J.W. Mercer, M. Kavanaugh, and F. DiGiano,
1988. Coping with Groundwater Contamination. J. Water
Pollution Control Federation 60(8): 1414-1428.
Campbell, M.D. and J.H. Lehr. 1973. Water Well Technol-
ogy. McGraw-Hill, New York, NY.
Cartwright, K. and J.M. Shafer. 1987. Selected Technical
Considerations for Data Collection and Intepretation-
Ground Water. In: National Water Quality Monitoring
and Assessment, National Academy Press, Washington,
DC, pp. 33-56.
de Vera, E.R. 1980. Samplers and Sampling Procedures for
Hazardous Waste Streams. EPA-600/2-80-018 (NTIS
PB80-135353).
Driscoll, F.G. 1986. Groundwater and Wells, 2nd ed. Johnson
Division, UOP, Inc., St. Paul, MN.
Englund, E. and A. Sparks. 1988 GEO-EAS (Geostatistical
Environmental Assessment Software) User's Guide. EPA/
600/4-88/033a (Guide NTIS PB89-151252; Software:
NTIS PB89-151245).
Everett, L.G. 1980. Ground Water Monitoring. Technology
Marketing Operation, General Electric Co., Schenectady,
NY.
Everett, L.G., L.G. Wilson, and E.W. Hoylman. 1984. Vadose
Zone Monitoring for Hazardous Waste Sites. Noyes Data
Corp., Park Ridge, NJ.
Ford, P.J., P.J. Tunna, and D.E. Seely. 1984. Characterization
of Hazardous Waste Sites - A Methods Manual, II, Avail-
able Sampling Methods, 2nd ed. EPA 600/4-84-076 (NTIS
PB85-521596). [The first edition was published in 1983
as EPA/600/4-83-040 (NTIS PB84-126920)].
20
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Gibb, J.P., R.M. Schuller, and R.A. Griffin. 1981. Procedures
for the Collection of Representative Water Quality Data
from Monitoring Wells. Cooperative Groundwater Re-
port 7. Illinois State Water Survey and Illinois State
Geological Survey, Champaign, IL.
Gillham, R.W., M.J.L. Robin, J.F. Barker and J.A. Cherry.
1983. Groundwater Monitoring and Sample Bias. API
Publication 4367. American Petroleum Institute, Wash-
ington, DC.
Guswa, J.H., WJ. Lyman, A.S. Donigan, Jr., T.Y.R. Lo, and
E.W. Shanahan. 1984. Groundwater Contamination and
Emergency Response Guide. Noyes Publication, Park
Ridge, NJ.
Hughes, J.P. and D.P. Lettenmaier. 1981. Data Requirements
for Kriging: Estimation and Network Design. Water Re-
sources Research 17(6): 1641-1650.
Keely, J.F., M.D. Piwoni, and J.T. Wilson. 1986. Evolving
Concepts of Subsurface Contaminant Transport. J. Water
Pollution Control Federation 58(5):349-357.
Keith, S.J., M.T. Frank, G. McCarty, and G. Mossman. 1983.
Dealing with the Problem of Obtaining Accurate Ground-
water Quality Analytical Results. In: Proc. Third Nat.
Symp. on Aquifer Restoration and Ground Water Moni-
toring, National Water Well Association, Dublin, OH,
pp. 272-283.
Kruseman, G.P. and N.A. de Ridder. 1976. Analysis and
Evaluation of Pumping Test Data. International Institute
for Land Reclamation and Improvement, Bulletin 11.
Wageningen, The Netherlands.
Larson, D. 1981. Materials Selection for Ground Water Moni-
toring. Presented at the National Water Well Association
Short Course entitled practical Considerations in the De-
sign and Installation of Monitoring Wells, Columbus,
OH, December 16-17.
Lawrence Berkeley Laboratory. 1977. Invitational Well-Test-
ing Symposium Proceedings. LBL-7027. Lawrence Ber-
keley Laboratory, Berkeley, CA.
Lawrence Berkeley Laboratory. 1978. Second Invitational
Well-Testing Symposium Proceedings. LBL-8883.
Lawrence Berkeley Laboratory, Berkeley, CA.
Lehr, J. H. 1988. An Irreverent View of Contaminant Disper-
sion. Ground Water Monitoring Review 8(4):4-6.
Mercer, J.W., S.D. Thomas, and B. Ross. 1982. Parameters
and Variables Appearing in Repository Siting Models.
NUREG/CR-3066. U.S. Nuclear Regulatory Commis-
sion, Washington, DC.
Mercer, M.W. and C.O. Morgan. 1981. Storage and Retrieval
of Ground-Water Data at the U.S. Geological Survey.
Ground Water 19(5):543-551.
Moore, I.E. 1979. Contribution of Ground-Water Modeling to
Planning. J. Hydrology 43:121-128.
Morrison, R.D. 1983. Ground Water Monitoring Technology,
Procedures, Equipment and Applications. TIMCO Manu-
facturing, Inc., Prairie du Sac, WI.
Nacht, SJ. 1983. Monitoring Sampling Protocol Consider-
ations. Ground Water Monitoring Review 3(3):23-29.
National Research Council (NRC). 1990. Ground Water Mod-
els: Scientific and Regulatory Applications. National
Academy Press, Washington, DC.
Olea, R.A. 1982. Optimization of the High Plains Aquifer
Observation Network, Kansas. Groundwater Series 7.
Kansas Geological Survey, Lawrence, KS.
Olea, R.A. 1984. Systematic Sampling of Spatial Functions.
Series on Spatial Analysis No. 7. Kansas Geological
Survey, Lawrence, KS.
Samper, F.J. and S.P. Neuman. 1985. Gcostatistical Analysis
of Hydrochemical Data from the Madrid Basin, Spain
(Abstract). Eos (Trans. Am. Geophysical Union)
66(46):905.
Scalf, M.R., J.F. McNabb, W.J. Dunlap, R.L. Cosby, and J.
Fryberger. 1981. Manual of Ground-Water Quality Sam-
pling Procedures. EPA/600/2-81/160, (NTIS PB82-
103045). Also published in NWWA/EPA Series, National
Water Well Association, Dublin OH.
Sisk, S.W. 1981. NEIC Manual for Groundwater/Subsurface
Investigations at Hazardous Waste Sites. EPA/330/9-81-
002 (NTIS PB82-103755).
USATHAMA. 1982. Sampling and Chemical Analysis Qual-
ity Assurance Program for U.S. Army Toxic and Hazard-
ous Materials Agency, Aberdeen Proving Ground, MD.
U.S. Environmental Protection Agency (EPA). 1986. RCRA
Ground Water Monitoring Technical Enforcement Guid-
ance Document. EPA OSWER-9950.1. Also published in
NWWA/EPA Series, National Water Well Association,
Dublin, OH.
van Geer, F.C. 1987. Applications of Kalman Filtering in the
Analysis and Design of Groundwater Monitoring Net-
works. TNO Institute of Applied Geoscience, Delft, The
Netherlands.
Virdee, T.S. and N.T. Kottegoda. 1984. A Brief Review of
Kriging and its Application to Optimal Interpolation and
Observation Well Selection. Journal des Sciences
Hydrologiques 29(4):367-387.
Wheatcraft, S.W. 1989. An Alternate View of Contaminant
Dispersion. Ground Water Monitoring Review 9(3): 11-
12.
21
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Chapter 3
Geologic Aspects of Site Remediation
James W. Mercer and Charles P. Spalding
This chapter addresses the geologic aspects of remedia-
tion: (1) What geologic factors are significant? (2) How are
geologic data collected? and (3) How are geologic data inter-
preted? To help answer these questions, this chapter includes
information on stratigraphy, lithology, structural geology, and
hydrogeology. However, this chapter does not include infor-
mation on basic geology, but the reader may consult any of
numerous textbooks on the subject. There is also a concise
review of basic geology in U.S. EPA (1987).
To support discussions of the geologic factors, means of
collecting geologic data are also included. See Chapter 4 for
specific, detailed information on wells. This chapter covers
soil and rock coring, as well as various surface and borehole
geophysical techniques. A case history on the Hyde Park
landfill concludes the chapter.
3.1 Stratigraphy
Stratigraphy is the study of the formation, composition,
sequence, and correlation of stratified rocks and unconsoli-
dated materials (e.g., clays, sands, silts, and gravels). Strati-
graphic data include formational designations, age, thickness,
areal extent, composition, sequence, and correlations. In a
stratigraphic investigation, aquifers and confining formations
are identified so that units likely to transport pollutants can be
delineated, and lateral changes in formations (facies changes)
are noted if present. In effect, the stratigraphy of a site defines
the geometry and framework of the ground-water flow sys-
tem. Therefore, knowledge of the stratigraphy is necessary in
order to identify pathways of chemical migration, to estimate
extent of migration, and to define the hydrogeologic framew-
ork.
The first step in conceptualizing a site is to study driller's
logs, well cuttings, and/or corings. While observations made
during drilling activities can provide additional information
such as drilling rates and water losses, the primary goal of
these observations is to characterize layers of like material.
This layering can be differentiated based on material type, but
a major consideration for characterization should be how well
the material transmits water. The primary differentiation should
be based on whether the material has properties of an aquifer
and readily transmits water or has properties of a confining
bed, prohibiting the movement of water.
Once the layering has been determined at each well, the
next task is to plot the wells at their relative locations to each
other and attempt to correlate the layers among the wells. This
correlation involves interpreting well-log data and requires
knowledge of geological processes. At some sites, the correla-
tion will be straightforward; at others, correlation may be
impossible, The ability to correlate also will depend on the
scale of the correlation. To understand the stratigraphic con-
trols of flow and chemical migration, only larger scale fea-
tures may need to be correlated. The completed correlation
results in a figure called a fence diagram (see Figure 3-1). As
shown in the figure, a fence diagram is composed of intersect-
ing geological cross-sections.
The elevations of where the layers connect can be con-
toured to form structural maps representing either the top or
bottom surface of various layers. Where dense immiscible
fluids are a concern, structural maps on top of confining layers
are valuable because such fluids will flow via gravity on top
of the confining layer toward the lower elevations. Structural
maps for adjacent units can be subtracted from each other to
yield thickness or isopach maps. An isopach map may be
used, for example, to show the overburden thickness of un-
consolidated material overlying bedrock. Once completed,
these maps, along with the fence diagram, will provide a
three-dimensional picture of the subsurface system through
which the ground water and chemicals are moving.
In addition to wells and well cuttings, other means to
obtain stratigraphic data include hand augers, split-spoon
samplers, shelby tubes, and rock-coring equipment. Hand
augers are useful, particularly in sandy materials, for examin-
ing soil profiles to shallow depths (a few meters) and for
installing monitoring devices. Many types of hand augers are
available, but all are limited to use in unconsolidated geologic
materials and tend to be impractical in dense clays or stony
matenals (Gillham, 1988).
A split-spoon sampler consists of a metal cylinder that is
split longitudinally and threaded on both ends. A cutting head
is threaded onto the lower end and a drill-rod attachment
threaded onto the upper end. The sampler is driven into the
formation at the bottom of an augered borehole, using a
drilling rig with a 140-pound weight (ASTM, 1990a). The
number of blows required to penetrate a soil is a function of
the compactability of the soil; thus, blow count can be used to
characterize soil types. When withdrawn and opened,, the
23
-------
Wilton
L Mann-3
R.O.
Smith-7
10 Miles
Canon Crk.
Figure 3-1. Sample fence diagram construction (from Compton, 1962).
sample is relatively undisturbed and shows the natural stratifi-
cation of the geologic material. Shelby tubes are thin-walled
metal tubes that are attached to drill rods and are driven into
the formation (Gillham, 1988). Samples can be sealed and
stored in the tubes and later extruded for examination. How-
ever, both shelby tubes and split-spoon samplers are limited to
sampling of unconsolidated materials.
When formations are too hard to be sampled by soil
sampling methods, core drilling can be used (ASTM, 1990b).
The simplest core barrel consists of a hollow steel tube with a
core catcher and a diamond or tungsten carbide core bit. Other
core barrels have a dual wall system with a floating inner
sleeve that remains stationary while the outer barrel rotates
and cuts the core. A wireline system is available that elimi-
nates pulling the drill pipe from the hole to recover each core
(Landau, 1987). In this system, the core material is retrieved
through the annulus of the drill rods.
Analysis of cores is performed both in the field and in the
laboratory. Laboratory analysis includes determination of po-
rosity; permeability; and saturation with respect to a specific
fluid component e.g., nonaqueous phase liquids (NAPL); and
lithology studies (Keelan, 1987). Field studies of cores in-
clude determination of rock quality designation (RQD), core
recovery rate, fracture nature and frequency, presence of
chemical odors, and general core lithology. RQD represents
the amount of core greater than 4 in. in length divided by the
length of core run attempted. This parameter is related to the
competence of the material core and the fracture density of the
core run, Because RQD often can be correlated to permeabili-
ty, it is useful in characterization studies. Often cores are
broken during transport so all fracture-related analyses should
be performed as soon as possible after the core has been
retrieved,
Coring also provides opportunities to monitor drilling
return fluids for both color changes related to lithology and
visual and olfactory evidence of contaminants. As coring
proceeds, net drilling fluid loss or gain to the cored formation
can be determined by maintaining an accurate balance of
drilling fluids used. Fluid losses to an interval may be the
result of fractures or solutioning within the rock matrix. As
rock of varying competence is encountered, drilling rate also
varies and for a given drilling system, drilling rate can be
characteristic of the material penetrated.
Because the conceptualization of site conditions is based
on roughly correlated parameters of subsurface and unseen
conditions, it is useful to construct a correlation chart of
selected parameters versus depth (see Figure 3-2). Additional
parameters that may have been included in this figure are
permeability, drilling rate per foot, water loss or gain, and
presence and type of contamination.
3.2 Lithology
Lithology is the study of the physical character and
composition of unconsolidated deposits or rocks. As dis-
cussed in the Handbook, it includes (1) mineralogy, (2) or-
ganic carbon content, (3) grain size, (4) grain shape, and (5)
packing. The first two items affect sorption, whereas the last
three items affect water storage and flow. Additionally, com-
paction and cementation will reduce permeability based on
primary porosity, whereas solution channels will increase
permeability (Levorsen, 1967).
The mineral composition of rocks and unconsolidated
deposits can be used to determine the chemical composition.
The chemical composition of the media affects chemical
transport in ground water via a variety of chemical reactions.
Such interactions primarily involve inorganics and include
24
-------
Depth (ft)
Driller
10 -
20 -
30 -
40 -
50 -
60 -
70 -
80 -
90 -
Interval
B
M.S.L
560-
550-
540-
530-
520-
510-
500-
490-
480.
Core
Box
Clay
Grain Structure
Size
I
I
0 50 100 SF FMCCel
Fracture
Count Per Ft
ROD
Pumping
Rate
(gpm)
02468 10 0 50 100 0 S 10
3A
3B
4A
4B
5A
SB
6A
6B
7A
7B
8A
8B
9A
~T7
o
Legend of "Structure" Indicators
Soft Sediment Deformation
Algae Beds (Stromatoliths)
Gypsum Band
O Favosite Coral (s)
•.'.• Vugs
Inclined Fracture
Figure 3-2. Correlation chart of hydrogeologic features (from GeoTrans, 1989).
/c Orchard
Member
25
-------
sorption, precipitation and dissolution, acid-base reactions,
complexation, and redox reactions. Examples of chemicals
that could be reduced to lower concentrations in ground water
through the formation of precipitates include arsenic (by
reaction with iron, aluminum, or calcium), lead (by reaction
with sulfide or carbonate), and silver (by reaction with sulfide
or chloride). Hydrolysis can lead to the precipitation of iron,
manganese, copper, chromium, and zinc contaminants. Oxi-
dation or reduction could favor the precipitation of chromium,
arsenic, and selenium.
The tendency of an organic chemical to sorb is directly
related to the fraction of total organic carbon content in terms
of grams of organic carbon per gram of soil. A typical value of
organic matter in mineral soils is 3.25 percent (Brady, 1974).
The amount of organic matter is approximately 1.9 times the
amount of organic carbon; therefore, a typical value for or-
ganic carbon content is 1.7 percent. However, data will vary
from site to site.
Although variation in sorption between different grain-
size fractions is mostly a reflection of their organic carbon
content, other factors such as surface area have an effect. In
general, the fine silt and clay fractions of soils have the
greatest tendency to sorb chemicals. Grain size also influ-
ences water storage and movement. The amount of soil in
each of various size groups is one of the major factors used in
analyzing and classifying a soil. Various agencies define soil
groups in slightly different ways (see Figure 3-3). In general,
coarser grained soil is more transmissive and has less storage
capacity than finer grained soil.
Grain shape also influences water storage and porosity
because grain shape affects the manner in which grains are
arranged. Highly angular and irregularly shaped, noncemented
grains tend to result in a greater porosity than smooth, regu-
larly shaped grains, although the difference may be slight
Grain-size analysis, conducted on samples from uncon-
solidated formations, yields the proportion of material in each
specified size range. Range distributions can be used to esti-
mate permeabilities, design monitoring wells, and enable
better stratigraphic interpretation. The results of a grain-size
analysis usually are plotted as shown in Figure 3-4. The sieve-
opening size retaining 90 percent of the soil is called the
effective particle size (D90%), whereas the sieve-opening size
retaining 50 percent is called the average particle size (D50%).
Uniform soils consist of grains of predominantly one size
yielding curves with steep slopes. Well-graded soils have
grains of many different sizes and, therefore, are characterized
by more gently sloping curves.
Soils composed of grains of nearly uniform particle size
have a larger porosity than a well-graded soil because, in the
well-graded soil, small particles occupy a portion of the
volume between the larger particles. In the vadose zone,
American
Society for
Testing and
Materials
American
Association of
State Highway
Officials
U.S.
Department of
Agriculture
Federal
Aviation
Administration
Corps of
Engineers,
Bureau of
Reclamation
Colloids'
Colloids'
Clay
Clay
Clay
Silt
Silt
Silt
Clay
Silt
Fine
Sand
Fine
Sand
Vefy Fine
F'ne Sand
Sand
Fines (Silt or Clay)"
Med
ium
San
Fine
Sand
Fine
Sand
Medium
Sand
Coarse
Sand
j {^ c/5
vu
Coarse
Sand
Medium
Sand
1 =
o $
Fine
Gravel
Fine
Gravel
Gravel
ll
Coarse
Gravel
Coarse
Gravel
Boulders
Cobbles
Gravel
§ "g Fine
Q ,55 Gravel
Sieve Sizes c\3§2 «>°° S •»
II I I
8 1 11 1§S S 8.3«
s §~ ^
CO ^f (O QQ Q
5S° S3'
1
a
$
Coarse
Gravel
Cobbles
*
8
8%8 §
Particle Size, mm.
* Colloids included in clay fraction in test reports.
** The LL and PI of "Silt" plot below the "A" line of the plasticity chart, Table 4,
and the LL and PI for "Clay" plot above the "A" line.
Figure 3-3. Soil-group size limits of ASTM, AASHO, USDA, FAA, Corps of Engineers, and USBR (from Portland Cement Associa-
tion, 1973).
26
-------
100
90
I 8°
| 70
60
50
40
30
20
10
I
Well Graded Soil
0.001
Figure 3-4.
0.01 0.1
Particle Diameter in MM
°40 Percent
0
10
20
30
40
50 \
60
70 '
80
90
100
Particle-size distribution for a uniform sand and a
well-graded soil (from Bouwer, 1978).
uniform soils develop a well-defined capillary fringe, whereas
well-graded soils tend to have a higher, but less distinct
capillary fringe.
In summary, much qualitative information concerning
properties that affect flow and transport can be gained from
lithology. At many hazardous waste sites, this type of infor-
mation may be all that is available in the early stages of field
study. Thus, it may be used to help guide subsequent phases of
the field work, such as well screen design. This type of
qualitative information may be very helpful in characterizing
vadose-zone properties, where hydrologic testing is more
difficult to conduct and interpret.
3.3 Structural Geology
Structural geology includes studying and mapping fea-
tures produced by movement after deposition. Structural fea-
tures include folds, faults, joints/fractures, and interconnected
voids (i.e., caves and lava tubes). Highly vesicular tops and
bottoms of basalt flows, for example, are often cited as
sources of significant permeability. Just as important to the
definition of structural features is the more rapid cooling and
more intense fracturing of the top and bottom of flows (Huntley,
1987). Deformed, inclined, or broken rock formations can
control topography, surface drainage, and ground-water re-
charge and flow. Joints and fractures are commonly major
avenues of water transport (preferential pathways) and usu-
ally occur in parallel sets.
Most fractures can be attributed to one of three causes
(Lcworsen, 1967). Some fractures format depth as a result of
an increase in rock volume from the folding and bending of
strata. Others are caused by the removal of overburden by
erosion in the zone of weathering. As sediments are unloaded
through erosion, the upper parts expand, and incipient weak-
nesses in the rocks become joints, fractures, and fissures.
Therefore, an increase of fracturing below an unconformity is
to be expected. Probably much of the initial solution channel-
ing through which surface waters percolate results from the
gradual increase in jointing and fracturing that accompanies
weathering. The third cause of fracturing is a reduction in the
volume of shales in the ground, due to diagenetic mineral
changes coupled with a loss of water during compaction.
Solution features, such as enlarged joints, sinkholes and
caves are common in limestone rocks and promote rapid
ground-water movement. Pertinent data on structural features
necessary to study and understand solution features include
type, compass orientation, dip direction and angle, and stratig-
raphy. Chapter 6 discusses the influence of fractured media on
ground-water flow and how it is characterized.
3.4 Hydrogeology
Hydrogeology concerns the relationship of the movement
of subsurface waters with geology, and ties stratigraphy,
! ithology, and structural geology to the theory of ground-
water hydraulics. The main goal in studying hydrogeology is
to determine directions and rates of ground-water flow. This
information is essential to any ground-water remediation or
ground-water monitoring program. Although this topic is
introduced in this section, it is discussed in more detail in
Chapter 4.
Hydrologic factors that are important to hydrogedogy
include surface drainage and surface water/ground-water rela-
tionships. Surface drainage information includes tributary
relationships, stream widths, depths, channel elevations, and
flow data. In a hydrogeologic investigation, the nearest per-
manent gaging station and period of record should be deter-
mined. A U.S. Geologic Survey (USGS) 7 1/2-minute
topographic map will show some of the necessary informat-
ion. Gaging stations and flow data can be identified and
obtained through USGS data bases. Streams either can receive
ground-water inflow or lose water by channel exfiltration. As
part of the investigation, hydrologic literature should be re-
viewed to determine if local streams are "gaining" or "losing."
Losing streams are common in areas of limestone bedrock and
those with arid climates and coarse-grained channel sub
strates. Potential ground-water recharge areas, sometimes in-
dicated by flat areas or depressions noted on the landscape,
also should be identified. Stereo-pair aerial photographs can
also be useful in these determinations (Ray, 1960). Irrigated
fields detected in aerial photographs suggest ground-water
recharge areas; swampy, wet areas suggest areas of ground-
water discharge.
Other important factors include aquifer delineation, back-
ground water quality, and depth to ground water. As used in
this Handbook, depth to ground water refers to the vertical
distance from the ground surface to the standing water level in
a well. In a confined aquifer, the depth to water represents a
point on a "piezometric" surface. The depths will limit the
equipment that can be used for purging and sampling. Infor-
mation should be collected to delineate aquifer type
(unconfined, confined, or perched); composition; boundaries;
hydraulic properties (permeability, porosity, transmissivity,
27
-------
etc.); and interconnection with other aquifers (direction of
leakage). These data are generally available through geologi-
cal survey publications.
Probable ground-water flow directions (both horizontal
and vertical) are determined by comparing the elevation of
water levels in different wells. The quality of ground and
surface water in an area should define, to a large extent,
potential uses. Knowledge of natural or background water
quality and water uses is required to assess contaminant
impacts. The quality of surface waters is usually available
from U.S. EPA, USGS, and state records. Ground-water data
will probably be limited for any given area, but may be
discussed through USGS Water Resources Division offices,
state geological surveys, and county health departments.
3.5 Hydrogeologic Investigations
Much of the data needed to understand site-specific
ground-water movement will be determined via hydrogeo-
logic investigations. The purpose of these investigations is to
determine flow directions, pathways and rates of ground-
water flow, potential receptors of ground water, potential
contaminants, and the extent of contamination in the subsur-
face. This information is required for selecting from altern-
ative remedial strategies, and it provides the framework for
design of a ground-water remedial program, if needed.
Some of the field methods used to obtain this information
include borehole exploration (including coring), mapping sur-
face features, and geophysical methods (both surface and
downhole techniques). Much of the information gained from
these methods will be helpful in interpreting the geology. For
ground-water flow information, additional field methods in-
clude (1) monitoring water elevations in wells and adjacent
surface waters, (2) performing aquifer tests (pumping and/or
slug tests), and (3) using special methods such as laboratory
analysis of cores and borehole flowmeters. For subsurface
chemistry, soil sample analysis must be performed, as well as
sampling and analysis of ground water. A typical monitoring
well for ground-water sampling is shown in Figure 3-5. If
nonaqueous phase liquid (NAPL) is present, any free product
thickness must be measured and sampling performed.
3.5.1 Geophysical Techniques
Geophysical techniques are used to better understand
subsurface conditions arid to delineate the extent of contami-
nation. Common surface techniques used at hazardous waste
sites include surface resistivity, electromagnetic surveys, seis-
mic reflection method, ground-penetrating radar, and magne-
tometer surveys (see Table 3-1).
In surface resistivity methods (Zohdy et al, 1974; Stewart
et al., 1983), the geologic materials act as part of a direct
current circuit. In general, there are two current electrodes and
two electrodes for measuring voltage differences. The electri-
cal potential measured between the electrodes depends on the
electrical properties of the geologic materials which, in turn,
depend upon the resistivity of the pore water and the amount
of pore water. Most soil and rock materials are highly resis-
tive, while water is highly conductive. Porosity and local
stratigraphy, therefore, can be deduced from the measure-
ments. Because of the concentrations of some solutes, con-
taminant plumes frequently appear as a highly conductive
layer. Resistivity methods, therefore, can be useful in identi-
fying and mapping certain plumes (Wish, 1983).
Electromagnetic instruments used in hydrogeologic in-
vestigations consist of a transmitter and receiver (Stewart,
1982). The transmitter produces an alternating magnetic field
that induces electrical currents within the ground. The in-
duced currents vary with the electrical conductivity of the
geological materials and alter the magnetic field of the trans-
mitter. This alteration is detected by a receiver. Generally,
these devices are carried by one person, and do not require the
installation of electrodes or geophones. They are likely to be
more cost-effective than resistivity methods because field
work can be completed more rapidly. They can be used to
detect changes in subsurface conductivity related to contami-
nant plumes or buried metallic waste such as drums (Green-
house and Slaine, 1983).
In surface seismic methods (Sverdrup, 1986), an impact
is made at a particular point on the ground surface using a
mechanical hammer or an explosive device. The resulting
sound waves are monitored by sensing devices (geophones)
positioned at various distances from the impact. The time of
arrival of the sonic waves depends on velocity and density
contrasts that occur as the wave passes through different
stratigraphic layers. By interpreting the sigml, the investiga-
tion determines the geologic layering in the area.
In ground-penetrating radar (Koemer et al., 1981), radio
waves are transmitted into the ground and the reflected waves
are monitored and analyzed. Reflections occur as a result of
geologic variations in porosity and water content. The method
is useful for determining stratigraphic variations and for locat-
ing buried objects such as steel drums.
Magnetometer surveys (Gilkeson et al., 1986) measure
the strength of the earth's magnetic field. A proton nuclear
magnetic resonance magnetometer is frequently used. One
person can rapidly perform a survey over a site of a few acres
by using this hand-held instrument. The surveyor sets up a
grid system and measures the magnetic field at each intersec-
tion of the grid. Areas with large amounts of buried metal,
such as steel drums, will have magnetic anomalies associated
with them. The strength of the anomaly will vary with the
amount and depth of the buried metal.
Borehole logging (Keys and MacCary, 1971; Keys, 1988)
includes a variety of methods involving lowering a tool into
the borehole (see Table 3-2). The tool measures the physical
properties of the geologic materials, or, alternatively, provides
an impulse or disturbance to the natural system, and measures
the response of the system to the disturbance. Common log-
ging tools include caliper, resistivity, neutron, gamma, and
sonic tools. Logging can proceed in both cased or uncased
boreholes, though most measurements can be made only
when the hole has not been cased. Most of the logging
methods are effective in distinguishing between sand and clay
and are, therefore, useful in locating zones of high permeabil-
ity (Kwader, 1986),
28
-------
Gas Vent Tube^
1/4" Gas Vent
Well Cap
Steel Protector Cap with Locks
Surveyor's Pin (flush mount)
Concrete Well Apron
Continuous Pour Concrete Cap
and Well Apron
(expanding cement)
Non-shrinking Cement
Potentiometric -_
sum/sedimentTra .-.••-".-••--
Figure 3-5. A typical monitoring well design (from GeoTrans, 1989).
Induction logging can be used to identify soil and rock
types, geologic correlations, soil and rock porosity, and pore
fluid conductivity. Resistivity logging is effective in identify-
ing soil and rock types, geologic correlations, soil and rock
porosity, pore fluid resistivity, and secondary permeability
such as the locations of fractures and solution openings.
Natural gamma logging can assist in positioning wells and
casings, by providing information on clay or shale content,
grain size, pore fluid resistivity, and soil and rock identifica-
tion. Gamma-gamma logging will help to position cementing
for the well casings and to determine total porosity or bulk
density. Neutron logs can provide estimates of moisture con-
tent above the water table, total porosity below the water
table, specific yield of confined aquifers, the location of the
water table outside the casing, chemical and physical proper-
ties of the water, and the rate of moisture infiltration. Tem-
perature logs help provide the chemical and physical
characteristics of the water source and movement of the
water in the well; and dilution, dispersion, and movement of
the waste.
Video cameras also have been developed that can be
lowered down a 4-in. (10-cm) diameter borehole. They can be
used for visual inspection and to provide a visual record of the
wall of the borehole. They are particularly useful for inspect-
ing the casing for corrosion, damage, or leaks, and also are
29
-------
Table 3-1. Summary of Surface Geophysical Methods
Surface Geophysical
Survey Method Applications
Advantages
Limitations
SEISMIC
REFRACTION
AND REFLECTION
Determines - Ground-water resource
lithological evaluations
changes in - Geotechnical profiling
subsurface - Subsurface stratigraphic
profiling including top of
bedrock
- Relatively easy accessibility
- High depth of penetration
dependent on source of
vibration
- Rapid areal coverage
Resolution can be obscured in
layered sequences
Susceptibility to noise from urban
development
Difficult penetration in cold weather
(depending on instrumentation)
Operation restricted during wet
weather
ELECTRICAL
RESISTIVITY
Delineates
subsurface
resistivity
contrasts due to
lithology, ground
water, and
changes in ground-
water qualify
ELECTROMAGNETIC
CONDUCTIVITY
Delineates
subsurface
conductivity
contrasts due to
changes in
ground-water
quality and
lithology
GROUND
PENETRATING
RADAR
- Depth to water table estimates
Subsurface stratigraphic profiling
Ground-water resource evaluations
High ionic strength contaminated
ground-water studies
Provides contin-
uous visual profile
of shallow sub-
surface objects,
structure, and
lithology
MAGNETICS
Detects presence
of buried metallic
objects
Subsurface stratigraphic profiling
Ground-water contamination studies
Landfill studies
Ground-water resource
evaluations
Locating buried utilities,
tanks, and drums
Locating buried objects
Delineation of bedrock
subsurface and structure
Delineation of karst features
Delineation of physical integrity of
manmade earthen structures
Location of buried ferrous
objects
Detection of boundaries
of landfills containing
ferrous objects
Location of iron-bearing rock
• Rapid areal coverage
• High depth of penetration
possible (400-800 ft)
• High mobility
• Results can be approximated
in the field
High mobility
Rapid resolution and data
interpretation
High accessibility
Effectiveness in analysis of very
high resistivity
Equipment readily accessible
Great areal coverage
High vertical resolution in
suitable terrain
Visual picture of data
High mobility
Data resolution possible in field
Rapid areal coverage
Susceptibility to natural and
artificial electrical interference
Limited use in wet weather
Limited utility in urban areas
Interpretation that assumes a
layered subsurface
Lateral heterogeneity
not easily accounted for
Data reduction less refined than with
resistivity
Use unsuitable in areas with surface
or subsurface power sources,
pipelines, utilities
Less vertical resolution than with
other methods
Limited use in wet weather
Limited depth of penetration
(a meter or less in wet, clayey
soils; up to 25 meters in dry, sandy
soils)
Accessibility limited due to bulkiness
of equipment and nature of survey
Interpretation of data qualitative
Limited use in wet weather
Detection dependent on size and
ferrous content of buried object
Difficult data resolution in urban areas
Limited use in wet weather
Data interpretation complicated in
areas of natural magnetic drift
Source: Modified after O'Brien and Gere (1988)
30
-------
Table 3-2. Summary of Borehole Log Applications
Required information on the Properties of
Rocks, Fluid Wells, or the Ground-Water System
Widely Available Logging Techniques That Might Be Utilized
Lithology and stratigraphic correlation of aquifers and
associated rocks
Total porosity or bulk density or gamma-gamma
Effective porosity or true resistivity
Clay or shale content
Permeability
Secondary permeability - fractures, solution openings
Specific yield of unconfined aquifers
Grain size
Location of water level or saturated zones
Moisture content
infiltration
Direction, velocity, and path of ground-water flow
Dispersion, dilution, and movement of waste
Source and movement of water in a well
Chemical and physical characteristics of water,
including salinity, temperature, density, and viscosity
Determining construction of existing wells, diameter and
position of casing, perforations, screens
Guide to screen setting
Cementing
Casing corrosion
Casing leaks and/or plugged screen
Electric, sonic, or caliper logs made in open holes. Nuclear logs made in
open or cased holes.
Calibrated sonic logs in open holes, calibrated neutron logs in open or
cased holes.
Calibratad long-normal resistivity logs.
Gamma logs.
No direct measurement by logging. Maybe related to porosity, injectivity,
sonic amplitude.
Caliper, sonic, or borehole televiewer or television logs.
Calibrated neutron logs.
Possible relation to formation factor derived from electric logs.
Electric, temperature, or fluid conductivity in open hole or inside casing. Neutron
or gamma-gamma logs in open hole or outside casing.
Calibrated neutron logs
Time-interval neutron logs under special circumstances or radioactive
tracers.
Single-well tracer techniques - point dilution and single-well pulse.
Multiwell tracer techniques.
Fluid conductivity and temperature logs gamma logs for some radioactive
wastes, fluid sampler.
infectivity profile. Flowmeter or tracer logging during pumping or injection.
Temperature logs
Calibrated fluid conductivity and temperature in the well Neutron chloride
logging outside casing. Multielectrode resistivity.
Gamma-gamma, caliper, collar, and perforation locator, borehole television.
All logs providing data on the lithology, water-bearing characteristics, and
correlation and thickness of aquifers.
Caliper, temperature, gamma-gamma. Acoustic for cement bond.
Under some conditions caliper or collar locator.
Tracer and flowmeter.
Source: Keys and MacCary (1971)
31
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used in uncased rock holes for locating fractures and fracture
zones (Gillharn, 1988).
3.5.2 Example-Hyde Park Landfill
This example, discussed in detail in Cohen et al. (1987),
concerns the Hyde Park landfill in Niagara Falls, New York
(see Figure 3-6). Ground-water studies were initiated at the
site in 1978 when a shallow tile drain and clay cover were
installed at the landfill. Remedial investigations (RI), required
by a settlement agreement, were conducted from 1982 to
1984. A major component of the RI was a drilling program
designed to determine the extent of chemical contamination in
the overburden and bedrock. Borings were cored and tested in
15-ft sections to the top of the Rochester Shale along 10
vectors radiating out from the landfill. Ground-water samples
were taken for analysis from those 15-ft sections that yielded
significant amounts of water. If chemicals were present above
specified levels, a new hole was drilled about 800 ft away
along the vector. Some of these holes were used as observa-
tion wells during aquifer tests prior to being grouted.
As a result of the drilling programs, the local geology is
fairly well known. Approximately 15 to 30 ft of waste at the
landfill are underlain by 0 to 10 ft of silty clay sediments. At
Hyde Park, the overburden lies unconformably on the Lockport
Dolomite. Undulations in the bedrock surface were carved by
previous glaciation. The Lockport Dolomite ranges in thick-
ness from 130 ft (200 ft southeast of the landfdl) to 65 ft at the
Niagara Gorge. The Lockport Dolomite overlies the Roches-
ter Shale and several lower units in a layer-cake sequence.
The hydrogeology of the Hyde Park area is unique be-
cause of the Niagara River Gorge and the human-induced
channels associated with a nearby pump storage reservoir (see
Figure 3-6). The Niagara Gorge (about 2,000 ft to the west),
the forebay canal (about 4,000 ft to the north), and the buried
conduits (about 3,000 ft to the east) control ground-water
movement in the Hyde Park area.
The ground-water system can be conceptualized as a
series of slightly dipping, permeable zones sandwiched be-
tween aquitards, all of which are bounded on three sides by
drains. Precipitation infiltrates the wastes and the low-perme-
ability overburden before recharging the highly fractured
upper layer of the Lockport Dolomite. Where glacial sedi-
ments are present beneath the landfill, downward ground-
water flow and chemical migration are retarded. In areas
Pumped-Storage
Reservoir
Hyde Park
Landfill
Buried
Conduits
Power
Canal
Robert Moses
Niagara
Power Plant
Niagara
River
-100
7,000 Feet
^"V Lockport Dolomite
0,000
Rochester Shale
Irondequoit
& Reynales Limestones
Sandstone & Shale
Whirlpool Sandstone
Queenston Shale
Figure 3-6. A generalized diagram showing the geologic formation and topographic features in the vicinity of the Hyde Park
landfiii (from Faust, 1985).
32
-------
where these sediments are thin and/or absent, ground water
and chemicals move freely into the underlying rock. In the
permeable bedrock zones, much of the ground water flows
laterally toward the three boundaries. Between these zones,
ground water moves slowly downward to the next lower
permeable layer. Pumping tests suggest an anisotropic system
where hydraulic conductivities are greatly affected by prefer-
ential flow along fractures (Figure 3-7). This conceptualization
is supported by the alignment and dip of joint systems ex-
pressed at nearby outcrops.
Analyses of ground-water samples taken during the vec-
tor well survey revealed that contamination had migrated
much further than previously thought. In fact, Hyde Park
chemicals were found in seeps emanating from the Lockport
Dolomite along the Niagara Gorge in July 1984. Dissolved
chemical and NAPL plumes in the overburden and in the
Lockport Dolomite were delineated during the RI as shown in
Figure 3-8. Although the areal extent of contamination has
been defined, the depth of chemical migration was unknown
because at many locations dissolved chemicals and NAPL
were observed all the way to the base of the Lockport Dolo-
mite.
The distribution of chemicals in the overburden reflects
the downward migration of contaminated surface runoff from
the Hyde Park landfill, which is elevated relative to surround-
ing properties. Lateral chemical transport through the
overburden has been limited because the potential for down-
ward flow to bedrock exceeds that for outward flow through
the low-permeability glacial sediments.
The contamination observed in the Lockport Dolomite
reflects variations in the directions of ground-water flow that
have occurred since waste disposal began at Hyde Park and, to
a lesser extent, at the dipping beds of the Lockport Dolomite.
Chemical analyses indicate the past migration of chemicals
through the upper Lockport Dolomite in all directions. Present
ground-water flow is primarily to the northwest, but the
southern and eastern areas of contamination suggest that at
one time ground water moved toward those areas. Ground-
water flow prior to the construction of the forebay canal and
buried conduits (from 1958 to 1962) was inferred to be toward
the southwest. Similarly, dewatering during the construction
of these conduits could have drawn contaminated ground
water toward the east. Chemicals have moved downward to
the base of the Lockport Dolomite by dissolution in ground
water and by dense NAPL flow.
The Hyde Park Stipulation requires several remedial ac-
tions, focusing on source control, overburden remedies, bed-
rock remedies, and control of seeps at the Niagara Gorge face.
The application of a series of numerical models of ground-
water flow and chemical transport facilitated these remedies.
The source control program is designed to reduce the
amount of chemicals migrating from the landfill into the
overburden and bedrock. This reduction will be achieved by a
synthetic cap to reduce recharge and by extraction wells to
remove chemicals. During the prototype phase of the pro-
gram, two large-diameter extraction wells will be installed in
the landfill. Exploratory boreholes were completed in the
landfill to characterize the overburden stratigraphy of the
landfill and to help determine stratigraphic controls on NAPL
movement. All exploratory boreholes were then converted to
NAPL monitoring wells. The success of the prototype extrac-
tion wells depends in part on the compatibility of the sandpack
with landfill materials. To test the selection of the well
sandpack, two sandpack materials were selected based on
known landfill constituents. If a reasonable amount of NAPL
can be removed with this method, an operational network of
six extraction wells will be installed.
The remedial program specified for the overburden is
designed to laterally contain the dissolved chemicals and
NAPL and to maximize collection of NAPL. Mobile NAPL
not removed from the overburden will tend to sink downward
to the bedrock and will be addressed by the bedrock remedy.
The overall approach of the program is to further define the
boundary of the overburden NAPL plume with a series of
borings and then install a tile drain to collect mobile NAPL.
The location and depth of the drain will be determined after
the overburden plume boundaries have been refined by a
series of 44 overburden borings around the landfill. As the
drain is installed, additional stratigraphic information will be
added as soil is removed. The performance criteria for the
overburden system are:
• An inward hydraulic gradient must be maintained toward
the drain or downward into the bedrock.
• There must be no expansion of the NAPL plume toward
the drain or downward.
Remedial systems planned for the Lockport Dolomite are
designed to contain both the NAPL and APL plumes. Specific
objectives of the bedrock remedial system are to contain
dissolved chemicals and NAPL within the NAPL plume,
contain dissolved chemicals in the area near the gorge face
that is designated the remediated APL plume, and eliminate
the seepage of chemicals at the gorge face. However, portions
of the APL plume will not be remediated. As with the source
control system, a prototype system will be implemented first
and later refined into an operational system. The system will
use extraction and injection wells to maximize the collection
of both dissolved chemicals and NAPL. The locations of
purge, injection, and monitoring wells, and a schematic cross-
section of the containment concept, are shown in Figures 3-9
and 3-10, respectively. The recirculation wells are added to
the NAPL plume containment system to speed up the recov-
ery of contaminants and to maintain higher water levels for
the flushing of chemicals in the upper bedrock.
All prototype bedrock extraction/injection wells and re-
lated Lockport monitoring wells will be completed in three
separate hydrogeologic zones. The separation of the Lockport
into three zones allows optimization of the remedial system
through better characterization, monitoring, and pumping
schemes for the selected zones.
The main performance criteria for the bedrock system is
the maintenance of an inward hydraulic gradient at the NAPL
plume boundary. In addition, the flux of certain chemicals to
33
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II UNIVERSITY | I DR
400'
I PENNSYLVANIA AVENUC
•IS
Legend
i Postulated Groundwater
' Drawdown Contour
Groundwater Observation Well
Bedrock Survey
Pumped Well
Figure 3-7. Postulated ground-water drawdown contours during Hyde Park landfill pump test (from Conestoga-Rovers & Associ-
ates Limited, 1984).
34
-------
I Bedrock APL Plume \
ADD Overburden Wells
• • Bedrock Wells
Figure 3-8. Boundaries of dissolved chemical (APL) and NAPL plumes of contaminated ground water emanating from the Hyde
Park landfill through the overburden and Lockport Dolomite (from Faust, 1985).
the Niagara River must be below specified limits. The interim
flux level for 2,3,7,8-TCDD is 0.5 g/yr. This level will be
modified based on a future study of TCDD in the Niagara
River and Lake Ontario.
Hyde Park is an excellent example of a remediation that
both allows for better site characterization and does not make
itself obsolete as more data become available. The remedies
described in the stipulation include extensive monitoring pro-
grams that both ensure that performance goals are achieved
and enhance the understanding of site hydrogeology. The
phased approach with initial prototype remedies allows for
better initial site characterization that will ultimately lead to
the optimal remediation approach. The program is not limited
to current technologies, but can be modified should new
innovations be found. This flexibility is important because of
the long cleanup times expected.
35
-------
• ' 0 500 1000 1500 Feet
I I I I
O Pwe IVe/te
Monitoring Wells
Injection Wells
Figure 3-9. Locations of purge, injection, and monitoring wells to be installed for the prototype Lockport Dolomite hydraulic
containment system at the Hyde Park site (from Faust, 1985).
36
-------
Hyde Park
Landfill
Before Remedial Pumping
Rochester r~^^^r_
Shale —
Figure 3-10. A conceptual cross-section of the Lockport Dolomite hydraulic containment system at the Hyde Park site
(from Faust, 1985).
37
-------
3.6 References
American Society for Testing and Materials (ASTM). 1990a.
Standard Method for Diamond Core Drilling for Site
Investigation. In: Annual Book of ASTM Standards, Vol.
04.08, D2113-83. ASTM, Philadelphia, PA.
American Society for Testing and Materials (ASTM). 1990b.
Standard Method for Penetration Test and Split-Barrel
Sampling of Soils. In: Annual Book of ASTM Standards,
Vol. 04.08, D 1586-84.
Bouwer, H. 1978. Ground-Water Hydrology. McGraw-Hill,
New York.
Brady, N.C. 1974. The Nature and Properties of Soils, 8th ed.
MacMillan, New York.
Cohen, R.M., R.R. Rabold, C.R. Faust, J.O. Rumbaugh III,
and J.R. Bridge. 1987. Investigation and Hydraulic Con-
tainment of Chemical Migration: Four Landfills in Niagara
Falls. Civil Engineering Practice 2(l):33-58.
Compton, R.R. 1962. Manual of Field Geology. John Wiley
& Sons, New York.
Conestoga-Rovers & Associates Limited. 1984. Requisite
Remedial Technology Study, Overburden & Bedrock,
Hyde Park Remedial Program. Prepared for Occidental
Chemical Corporation, Ref No. 1069.
Faust C.R. 1985. Affidavit re: Hyde Park. U.S., N.Y. v.
Hooker Chemicals and Plastics Corp. et al, Civil Action
No. 79-989.
GeoTrans. 1989. Progress Report - Hydrogeological Charac-
terization of the Bedrock Near the S-Area Landfill (Niagara
Falls, NY) in Support of Requisite Remedial Technology
(RRT) Evaluation. Geo Trans, Inc., Hemdon, VA.
Gilkeson, R.H., P.C. Heigold, and D.E. Laymen. 1986. Prac-
tical Application of Theoretical Models to Magnetometer
Surveys of Hazardous Waste Disposal Sites—A Case
History. Ground Water Monitoring Review 6(1):54-61.
Gillham, R.W. 1988. Glossary of Ground Water Monitoring
Terms. Water Well Journal 42(5):67-71.
Greenhouse, J.P. and D.J. Slaine. 1983. The Use of Recon-
naissance Electromagnetic Methods to Map Contaminant
Migration. Ground Water Monitoring Review 3 (2): 47-
59.
Huntley, D. 1987. Some Fundamentals of Hydrogeology, 5th
ed. In: Subsurface Geology, L.W. LeRoy, D.O. LeRoy,
S.D. Schwochow, and J.W. Raese (eds.), Colorado School
of Mines, Golden, CO, pp. 746-755.
Keelan, O.K. 1987. Core Analysis. In: Subsurface Geology,
5th d., L.W. LeRoy, D.O. LeRoy, S.D. Schwochow, and
J.W. Raese (eds.), Colorado School of Mines, Golden,
CO, pp. 35-47.
Keys, W.S. and L.M. MacCay. 1971. Application of Bore-
hole Geophysics to Water-Resources Investigations. U.S.
Geological Survey Techniques of Water-Resources In-
vestigations TWI-2E1.
Keys, W.S. 1988. Borehole Geophysics Applied to Ground-
Water Investigations. U.S. Geological Survey Open-File
Report 87-539,303 pp. [Published in 1989 with the same
title by National Water Well Association, Dublin, OH.]
Koemer, R.M., A.E. Lord, Jr., and J.J. Bowders. 1981. Utili-
zation and Assessment of a Pulsed RF System to Monitor
Subsurface Liquids. In: National Conference on Manage-
ment of Uncontrolled Hazardous Waste Sites, Hazardous
Materials Control Research Institute, Silver Spring, LD,
pp. 165-170.
Kwader, T. 1986. The Use of Geophysical Logs for Determin-
ing Formation Water Quality. Ground Water 24:11-15.
Landau, H.L. 1987. Coring Techniques and Applications. In:
Subsurface Geology, 5th cd., L.W. LeRoy, D.O. LeRoy,
S.D. Schwochow, and J.W. Raese, (eds.), Colorado School
of Mines, Golden, CO, pp. 395-398.
Levorsen, A.I. 1967. The Reservoir Pore Space. In: Geology
of Petroleum, 2nd ed, W.H. Freeman and Company, San
Francisco, CA, Chapter 4, pp. 115, 119, 120.
O'Brien and Gere Engineers, Inc. 1988. Hazardous Waste
Site Remediation. Van Nostrand Reinhold, New York,
422 pp.
Portland Cement Association. 1973. PCA Soil Primer. Engi-
neering Bulletin EB007.045, Portland Cement Associa-
tion, Skokie, IL, 39 pp.
Ray, R.G. 1960. Aerial Photographs in Geologic Interpreta-
tion and Mapping. U.S. Geological Survey Professional
Paper 373,230 pp.
Stewart, M., M. Layton, and T. Lizanec. 1983. Application of
Surface Resistivity Surveys to Regional Hydrogeologic
Reconnaissance. Ground Water 21:42-48.
Stewart, N.T. 1982. Evaluation of Electromagnetic Methods
for Rapid Mapping of Salt-Water Interfaces in Coastal
Aquifers. Ground Water 20:538-545.
Sverdrup, K.A. 1986. Shallow Seismic Refraction Survey of
Near-Surface Ground Water Flow. Ground Water Moni-
toring Review 6(1):80-83.
Urish, D.W. 1983. The Practical Application of Surface Elec-
trical Resistivity to Detection of Ground-Water Pollution.
Ground Water 21:144-152.
U.S. Environmental Protection Agency (EPA). 1987. Hand-
book Ground Water. EPA/625/6-87/016, 212 pp.
Zohdy, A.A.R, G.P. Eaton, and D.R. Mabey. 1974. Applica-
tion of Surface Geophysics to Ground-Water Investiga-
tions. U.S. Geological Survey Techniques of
Water-Resources Investigations TWI-2D1.
38
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Chapter 4
Characterization of Water Movement in the Saturated Zone
James W. Mercer and Charles P. Spalding
Advection is the primary transport mechanism for con-
servative chemicals and for many nonconservative chemicals.
It is the controlling process for chemicals moving away from
a source area (e.g., a landfill or a spill) and for removing
chemicals from the subsurface (e.g., pump-and-treat systems).
Therefore, understanding advection is important to both site
characterization and remediation.
An understanding of the factors that control ground-water
movement is needed to understand advection. This chapter
reviews concepts needed to determine and understand ground-
water flow. This review is followed by a discussion of field
techniques used to obtain the data needed to characterize
ground-water flow. As important as it is to collect data, it is
just as important to analyze and interpret the data. Therefore,
this chapter also discusses different analysis techniques and
ground-water remedial actions. Finally, an example ties the
discussion together and illustrates the important points of the
chapter.
Both general data requirements and characterization tech-
niques are presented throughout this chapter. Each application
of these techniques is unique and site specific. No subsurface
characterization tool provides perfect information; several
techniques (e.g., geophysical and geochemical) should be
combined, such that different types of data support the same
conclusion. Because the field work is completed in phases,
remediation decisions often involve some uncertainty; there-
fore, the importance of monitoring is stressed.
4.1 Review of Concepts
There are numerous books that characterize and present
the principles and concepts of ground-water hydrology (e.g.,
Bear, 1979; Bouwer, 1978; Davis and DeWiest, 1966; DeWiest,
1969; Domemco, 1972; Freeze and Cherry, 1979; Todd, 1980
and Walton, 1970). Other general references have been pub-
lished by the U.S. Environmental Protection Agency (e.g.,
U.S. EPA, 1987). This section specifically discusses contamin-
ant hydrology and will not cover many of the general topics
included in these references.
At hazardous waste sites, the following questions need to
be addressed with respect to ground-water hydrology: (1)
where is the water coming from? (2) where is the water going?
and (3) what are the rates of movement? Answering these
questions requires information on the local water balance, the
transmissive properties of the media, and the hydraulic head
distribution.
Hydraulic head is rhe elevation to which water rises in a
well that is open to the surface (Figure 4-1). It is composed of
two parts: (1) the pressure head that produces the column of
water above the open interval; and (2) the elevation head,
which is the elevation of the open interval relative to a datum,
usually mean sea level. Depth to water normally is measured
from a reference point (e.g., top of the casing) that has been
surveyed. This information is used to compute water-level
Soil-Water Systems
Saturated Unsaturated
Piezometer
Tensiometer
Porous Cup
11 I 1
Figure 4-1. A diagram of the relationships between hydraulic
head, H, pressure head, h, and gravitational head,
Z. The pressure head is measured from the level
of termination of the piezometer or tensiometer in
the soil to the water level in the manometer and is
negative in the unsaturated soil.
39
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elevations. Although depth to water is useful to know, without
converting it to a water-level elevation (i.e., hydraulic head),
directions and rates of ground-water movement cannot be
determined.
Hydraulic head data m often displayed in two dimen-
sions as a potentiometric surface map (Figure 4-2). Such a
map represents the elevation to which water would rise in an
open well placed in the interval of interest. It is analogous to a
topographic map with the direction of water flow from higher
to lower elevations and generally running perpendicular to the
contours. However, ground-water flow directions may di-
verge from the direction predicted by potentiometric contours
when the aquifer is anisotropic (hydraulic conductivity is not
the same in all directions). Fetter (1981) describes techniques
for determining the direction of ground-water flow in aniso-
tropic aquifers. Again, using the analogy of the topographic
map, behavior of ground-water flow is similar to how surface
Ground-water
Flow Direction
422'
420' ,.
418' -
416' .-•'
Legend
420
P3
Potentiometric Surface (NGVD)
Piezometer Location
BMx437.24 Bench Mark
TBM
4/7.37
Temporary Bench Mark
Property Line
200
Feet
400
Figure 4.2. Potentiometric surface map (from EPA, 1988).
40
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water runoff occurs via overland flow. Different subsurface
units or intervals may have different potentiometric surface
maps. The uppermost potentiometric surface map, which is in
contact with the atmosphere through the vadose zone (Chapter
5), is the water table.
Although displayed on a two-dimensional surface, the
hydraulic head distribution is generally a three-dimensional
phenomena that is, hydraulic head varies vertically as well as
areally. To determine the vertical distribution of hydraulic
head, wells must be drilled in the same vicinity, but must be
open to different depths (elevations). If hydraulic head in-
creases with increasing depth, ground-water flow is upward;
in general, this results in an area of discharge. If hydraulic
head decreases with increasing depth, ground-water flow is
downward this is an area of ground-water recharge.
Often the stratigraphy supports multiple aquifers that are
separated by confining beds. In these cases, the aquifers are
dominated by horizontal flow and the confining layers are
dominated by vertical flow, i.e., leakage between adjacent
aquifers. At hazardous waste sites, it is important to determine
how many aquifers are contaminated. As part of this determi-
nation, the direction of leakage and the direction of flow in the
affected aquifers must be assessed. It is possible that flow
direction in one aquifer could differ from flow in an adjacent
aquifer. The difference in hydraulic head over a given dis-
tance is known as the hydraulic gradient. Hydraulic gradients
must be known to determine rates and directions of ground-
water movement.
Often, topographic highs are recharge areas and topo-
graphic lows are discharge areas. For this reason, surface
water bodies (such as lakes, rivers, springs, and seeps) are
often surface expressions of the water table. Therefore, these
surface water bodies are useful for inferring watertable eleva-
tion data where no wells exist.
As indicated, ground water generally flows from poten-
tiometric highs to potentiometric lows, following a trace that
is perpendicular to the potentiometric contours. This trace is
sometimes referred to as a flow line. Unlike surface water,
however, ground-water flow is resisted by the rock and soil
through which it flows. This resistance is quantified by the
transmissive properties of the media. As these transmissive
properties vary at different locations in the aquifer and in
different directions from a given point, they cause the flow
lines to change directions such that they may no longer be
perpendicular to the apparent potentiometric contours. There-
fore, in addition to hydraulic gradients, the transmissive prop-
erties of the media must be known in order to determine rates
and directions of ground-water flow.
The transmissive properties of the media have been given
different but related terms, including intrinsic permeability,
hydraulic conductivity, and transmissivity. Intrinsic perme-
ability is a property of the porous medium and has dimensions
of length squared. It is a measure of the resistance to fluid
flow through the medium; the greater the permeability, the
less the resistance. Hydraulic conductivity is defined as the
volume of water that will move in unit time under a unit
hydraulic gradient through a unit area measured at right
angles to the direction of flow. It is a property of the fluid and
medium with dimensions of length per time. It is equal to the
product of intrinsic permeability, density of water, and the
gravitational acceleration constant divided by the dynamic
viscosity of water. Finally, transmissivity is the rate of water
flow through a vertical strip of aquifer one unit wide, extend-
ing the full saturated thickness of the aquifer, under a unit
hydraulic gradient. It is equal to the product of hydraulic
conductivity and the aquifer thickness. Consequently, it has
dimensions of length squared per time.
All these properties can vary spatially and directionally at
a given point. If the medium is homogeneous, the transmis-
sive properties do not vary spatially. If the medium is isotro-
pic, they do not vary when measured in different directions
from a given point. Most geologic materials are heteroge-
neous and anisotropic.
The final information that is needed to answer the ques-
tions about ground-water hydrology that were posed earlier
concerns the local water balance. At hazardous waste sites, it
is generally not possible to accurately quantify the local water
balance, primarily because of data limitations. One of the
main goals at a hazardous waste site investigation is to define
the extent of contamination. Consequently, monitoring wells
are clustered near potential sources. Ground-water hydrolo-
gists look at the bigger picture to determine what hydraulic
boundaries control or influence flow at the site. Data covering
the larger area are rarely available for hazardous waste sites.
Regardless, it is important to attempt the mass balance and to
estimate what regional factors control the local flow system.
This exercise, while not highly quantitative, will provide a
valuable qualitative understanding of the flow system control-
ling contaminant migration.
4.2 Field Techniques
Ground water is generally below the land surface and,
therefore, difficult to observe. One of the most effective
techniques for observing ground water is to use point mea-
surements made in wells. Wells must be designed, drilled, and
developed in order to measure water levels and to take water
quality samples. Tests are conducted to determine transmis-
sive and storage properties. The following section discusses
methods used to drill wells, measure water levels, and deter-
mine subsurface properties.
4.2.1 Drilling Techniques
Table 4-1 summarizes the advantages and disadvantages
of various drilling methods used for monitoring well construc-
tion. In shallow unconsolidated deposits, a hollow stem con-
tinuous flight auger is the preferred method. The use of
hollow stem augers (Figure 4-3) requires no fluid in the
borehole and allows for installation of the casing and screens
prior to removal of the augers, thereby eliminating problems
associated with caving of the borehole. However, it may be
difficult to seal the annular space in wells constructed in this
manner, and other construction techniques may be more suit-
able. In situations where borehole caving is not a problem, the
use of solid stem or bucket augers is equally suitable. Unfortu-
nately, the use of augers becomes impractical when drilling
41
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Table 4-1. Auger, Rotary, and Cable-Tool Drilling Techniques-Advantages and Disadvantages
for Construction of Monitoring Wells
Type Advantages Disadvantages
Auger .Minimal damage to aquifer
.No drilling fluids required
.Auger flights act as temporary casing, stabilizing hole for
well construction
.Good technique for unconsolidated deposits
.Continuous core can be collected by wireline method
Rotary .Quick and efficient method
.Excellent for large and small diameter holes
.No depth limitations
.Can be used in consolidated and unconsolidated deposits
.Continuous core can be collected by wireline method
• Cannot be used in consolidated deposits
• Limited to wells less than 150 feet in depth
• May have to abandon holes if boulders are encountered
.Required drilling fluids which alter water chemistry
.Results in a mud cake on the borehole wall, requiring
additional well devopment, and potentially causing
changes in chemistry
.Loss of circulation can develop in fractured and high-
permeability material
Cable tool .No limitation on well depth
.Limited amount of drilling fluid required
.Can be used in both consolidated and unconsolidated
deposits
.Can be used in areas where lost circulation is a problem
.Good lithologic control
.Effective technique in boulder environments
.Limited rigs and experienced personnel available
.Slow and ineffcient
.Difficult to collect core
From GeoTrans, 1989
deeper wells (100 to 150 ft) or when hard unconsolidated
deposits are encountered. Thick clay deposits that tend to bind
augers also may make the use of augers impractical. When
drilling beneath the water table where cross-contamination
between water-bearing strata is considered problematic, au-
gers may not be the optimum technique. If auger techniques
are used, it may not be possible to prevent fluid flow in the
borehole between formations.
When drilling in deeper consolidated deposits, air rotary
drilling (Figure 4-3) is frequently the preferred method be-
cause no drilling fluids are employed. However, oil from air
compressors may contaminate the borehole, and special filters
are required to minimize this effect. In some cases, drillers
may use foams to help lift cuttings to the surface and increase
the speed of drilling. Caving of unconsolidated material over-
lying consolidated material can frequently limit the use of air
rotary drilling. However, some air rotary rigs are equipped
with casing hammers that can drive a casing as drilling
proceeds, similar to cable tool drilling techniques (see discus-
sion below). Mud rotary techniques also can be used to drill
through unconsolidated material, a casing can be set to hold
these deposits open, and the hole can be continued with air
rotary.
Cable tool drilling methods (Figure 4-3) may be used for
constructing monitoring wells. However, cable tool drilling
through unconsolidated material, particularly below the water
table, will probably require the simultaneous driving of a
casing to prevent caving. Because casing driven in this man-
ner may seal strata through which it is driven, this method
may be used at sites when cross-contamination of water
bearing zones could be a problem. Completing a well cased
during drilling will probably require that the casing be pulled
back to expose the formation before setting the screens. An
advantage of cable tool drilling is that it can be used to drill to
great depths, although a minimum borehole diameter of 3 in.
is required. Another advantage is that it can penetrate through
consolidated material, although frequently at a slow pace.
During drilling for any ground-water contamination in-
vestigation, precautions must be taken to prevent cross-con-
tamination of boreholes. Thoroughly cleaning the drilling rig
and tools initially and after each borehole is drilled are ex-
amples of specific precautions that should be taken. No uni-
form procedure has been developed for all sites, but a soap
wash followed by solvent and distilled water rinse is com-
monly used. Proper drilling plans also can minimize potential
cross-contamination. If possible, drilling should progress from
the least to most contaminated areas (Sisk, 1981).
Upon completion, the monitoring well must be devel-
oped. Any contamination or formation damage from well
drilling and any fines from the natural formation must be
removed to provide a particulate-free discharge. A variety of
techniques are available to remove such contamination and
develop a well (Table 4-2). To be effective, all these tech-
niques require reversals or surges in flow to avoid bridging by
particles, which is common when flow is continuous in one
direction. These reversals or surges can be created by using
surge blocks, air lifts, bailers, or pumps (see Scalf et al.,
1981). Natural formation water should be used; use of other
water is not recommended. The discharge from the well
should be continuously monitored and development should be
continued until the discharge is particulate-free. Ideally, the
well should be developed so as to minimize the creation of
water requiring disposal.
42
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Air, Water or
Drilling Fluid
J
Auger
Flight
Cable •
I
X
Drill Stem''
Drill Bit,
n
Hollow-Stem Auger
Direct Rotary
Cable Tool
Figure 4-3. A conceptual comparison of the hollow-stem auger, the direct-rotary, and the cable-tool drilling methods (from
GeoTrans, 1989).
Table 4-2. Well Development Techniques-Advantages and Disadvantages
Technique Advantages
Overpumping
Backwashing
Mechanical surging
High velocity jetting
.Minimal time and effort required
.No new fluids introduced
.Remove fluids introduced during drilling
.Effectively rearranges filter pack
.Breaks down bridging in filter pack
.No new fluids introduced
.Effectively rearranges filter pack
.Greater suction action and surging than
backwashing
.Breaks down bridging in filter pack
.No new fluids introduced
.Effectively rearranges filter pack
.Breaks down bridging in filter pack
.Effectively removes the mud cake around
screen
Disadvantages
.Does not effectively remove fine-granted sediments
.Can leave the lower portion of large screen intervals
undeveloped
.Can result in a large volume of water to be contained and
disposed.
.Tends to push fine-grained sediments into filter pack
.Potential for air entrapment if air is used
.Unless combined with pumping or bailing, does not
remove drilling fluids
. Tends to push fine-grained sediments into filter pack
.Unless combined with pumping or bailing, does not
remove drilling fluids
.Foreign water and contaminants introduced
.Air blockage can develop with air jetting
.Air can change water chemistry and biology (iron bacteria)
near well
.Unless combined with pumping or bailing, does not
remove drilling fluids
From GeoTrans, 1989
43
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A variety of materials are available for use in casing,
screenings and other structural and sampling components of
monitoring wells (Table 4-3). Well materials must have suffi-
cient strength to ensure the structural integrity of the well
during installation and during protracted periods of monitor-
ing. The materials should sufficiently resist deterioration that
may result from long-term exposure to natural chemical or
pollutant constituents in the ground water at each site. The
materials also must be selected to minimize their interference
with the measurement of specific constituents. The most
commonly used materials are mild steel, stainless steel, poly-
vinyl chloride (PVC), polypropylene, polyethylene, and
Teflon®. These materials have substantially different proper-
ties relative to strength, corrosion resistance, interference with
specific constituent measurements expense, and availability.
Consequently, materials should be selected only after consid-
eration of all pertinent, site-specific factors such as well
installation method, depth, geochemical environment, and
probable contaminants to be monitored. Larson (1981) and
Barcelona et al. (1983) have summarized the chemical resis-
tance of various casings and well materials to differing envi-
ronments. These topics also are discussed in more detail in
subsequent chapters of this Handbook.
There are three basic categories of monitoring well de-
signs that are used to monitor vertical distribution of contami-
nants at a specific location (Figure 4-4). The first type of
nested-sampler design consists of a series of multiple-port
samplers installed in a single borehole. The sampling ports are
isolated from each other by inflatable packers or by other
annular seals. In some systems, a special tool is lowered into
the well to open ports at the specific location when a water
level or water quality sample is desired. In others, different
plastic (such as nylon) tubings are used for sampling each
zone where a vacuum is used to bring the sample to the
Table 4-3. Well Casing and Screen Material—Advantages and Disadvantages in Monitoring Wells
Type Advantages Disadvantages
Fluorinated ethylene
propylene (FEP)
Good chemical resistance to volatile organics
Good chemical resistance to corrosive
environments
Lower strength than steel and iron
Polytetrafluoroethylene
(PTFE) or Teflon®
Lightweight
High-impact strength
Resistant to most chemicals
Weaker than most plastic material
Polyvinylchloride
(PVC)
Polyethylene
Polypropylene
Kynar
Stainless steel
Cast iron and low-carbon steel
Galvanized steel
Lightweight
Resistant to weak alkalis, alcohols, aliphatic
hydrocarbons, and oils
Moderately resistant to strong acids and alkalis
Lightweight
Lightweight
Resistant to mineral acids
Moderately resistant to alkalis, alcohols, ketones,
and esters
High strength
Resistant to most chemicals and solvents
High strength
Good chemical resistance to volatile organics
High strength
High strength
Weaker than steel and iron
More reactive than PTFE
Deteriorates when in contact with ketones,
esters, and aromatic hydrocarbons
Low strength
More reactive than PTFE, but less reactive
than PVC
Not commonly available
Low strength
Deteriorates when in contact with oxidizing
acids, aliphatic hydrocarbons, and aromatic
hydrocarbons
More reactive than PTFE, but less reactive
than PVC
Not commonly available
Poor chemical resistance to ketones, acetone
Not commonly available
May be a source of chromium in low pH
environments
May catalyze some organic reactions
Rusts easily, providing highly sorptive surface
for many metals
Deteriorates in corrosive environments
May be a source of zinc
If coating is scratched, will rust, providing a
highly sorptive surface for many metals
From GeoTrans, 1989
44
-------
Multiple Port
Samplers
Open
Borehole
or \
Filter Pack
Packer or
Annular Seal
Sampling
Ports
Multiple Wells
Single Borehole
Multiple Wells
Multiple Boreholes
I
!
Borehole
Wall
• Annular Seals •
Screens
Figure 4-4. A conceptual comparison of three multilevel sampling designs (from GeoTrans, 1989).
surface. However, for deep wells and volatile organic chemi-
cals, the vacuum may result in unacceptable chemical losses
from volatilization. The second configuration for nested-sam-
plers consists of multiple well stings installed in one large
borehole (Figure 4-4). Individual zones are isolated from each
other using a low permeability material. Seals between zones
may be difficult to obtain and maintain.
Finally, the third type of nested-sampler design consists
of drilling a separate borehole for each monitoring well (Fig-
ure 4-4). This system is superior to the two previous systems
because the potential for cross- contamination from faulty
seals is minimized, and smaller diameter holes can be drilled,
thereby reducing the volume of water that needs to be pumped
prior to sampling. The additional costs associated with drill-
ing multiple boreholes often is offset by technical problems
associated with the installation of the two previous systems.
Use of multiple piezometers and ports in a single borehole
should be avoided according to U.S. EPA (1986), because the
potential for erroneous data is increased. (A piezometer is a
small-diameter well open to a point in the subsurface.) Table
4-4 summarizes the advantages and disadvantages of these
three multilevel sampling designs.
4.2.2 Methods to Measure Hydraulic Head
There are a number of ways to measure hydraulic head in
the saturated or vadose zones (see Table 4-5). For conve-
nience, both zones are discussed in the following section. The
accuracy of depth-to-water measurements is discussed in a
Superfund ground-water issue paper (Thornhill, 1989). When
comparing various methods of measurements, as indicated in
Table 4-5, the steel tape method is the most precise. Although
less precise, the air line method is useful in pumped wells
where water turbulence exists. Pressure transducers can be
used in either the saturated or vadose zones. They are useful
for making frequent measurements, such as during a slug test.
In a saturated zone, the hydraulic head, H, is measured at
a point using a piezometer (see Figure 4- 1) and is defined as
the elevation (pressure head) at which the water surface stands
in an open piezometer tube terminated at a given point in the
porous medium. Hydraulic head is a combination of pressure
head and elevation head (distance of the measuring point
above a reference level [datum]). The reference level chosen
for measurement of H is arbitrary. The hydraulic head is a
potential function, the potential energy per unit weight of the
ground water.
45
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Table 4-4. Multilevel Monitoring Well Design-Advantages and Disadvantages in Monitoring Wells
Type Advantages Disadvantages
Multiple-port sampler
.Large number of sampling zones per borehole
.Smaller volume of water required for purging than
nested sampler/single borehole and multiple boreholes
.Lower drilling costs than nested sampler/multiple
boreholes
.Potential for cross-contamination among ports
.Potential sampling ports becoming plugged
.Special sampling tools required
Nested sampler/
single borehole
.Lower drilling costs than multiple boreholes
.Low potential for screens becoming plugged
.Potential for cross-contamination among screen
intervals
.Number of sampling intervals limited to three or
four
.Larger volume of water required for purging
than multiple-port campier or nested sampler/
multiple boreholes
.Higher installation costs
Nested sampler/
multiple boreholes
.Potential for cross-contamination minimized
, Voliume of water required for purging smaller than
nested sampler/single borehole
.Low installation costs
.Low potential for screens becoming plugged
.Higher drilling costs
From GeoTrans, 1989
Table 4-5. Summary of Methods to Measure Hydraulic Head
Method Application
Reference
Steel tape Saturated zone. Most precise method. Noncontinuous measurements. Slow
Electric probe Saturated zone. Frequent measurements possible. Simple to use.
Adequate precision
Garber and Koopman (1968)
Driscoll (1986)
Air line
saturated zone. Continuous measurements. Useful for pumping tests.
Limited accuracy
Driscoll (1986)
Mechanical float Saturated zone. Continuous measurements. Useful for long-term measurements. USGS (1977)
recorder Permanent record can be delicate
Pressure saturated or vadose zone. Continuous or frequent measurements. Rapid response
transducer to changing pressure. Permanent record. Expensive
Acoustic sounder Saturated zone. Fast; permanent record. Imprecise
Tensiometry Saturated or vadose zone. Laboratory or field method. Useful range is 0 to 0.85
bars capillary pressure. Direct measurement. A widely used method
Electrical Vadose zone. Laboratory or field method. Useful range is 0 to 15 bars capillary
resistivity pressure, indirect measurement. Prone to variable and erratic readings
Thermocouple Vadose zone. Laboratory or field method. Useful range 10 to 70 bars capillary
psychrometry pressure, interference from dissolved solutes likely in calcium-rich waste
Thermal Vadose zone. Laboratory or field method. Useful range O to 2.0 bars capillary
diffusivity pressure, indirect measurement
Copyright® 1989 Electric Power Research Institute. EPRI EN-6637. Techniques to Develop Data for Hydrogeochemical Models. Reprinted with
permission.
Modified from Thompson et a/., 7989
Gerber and Koopman (1986)
Davis and DeWiest (1966)
Cassei and Klute (1986):
Stannard (1986)
Campbell and Gee (1986);
Rehm et al. (1987)
Rawlins and Campbell (1986)
Phene and Beale (1976)
46
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These same concepts of hydraulic head, pressure head,
and gravitational (or elevation) head may be applied to the
vadose zone (Chapter 5). A common device used to measure
the hydraulic head in the vadose zone is a tensiometer. It is
terminated in the soil by a porous cup permeable to water, but
impermeable to air, when the pores of the cup are filled with
water. The porous cup is necessary to establish hydraulic
contact between the water in the tensiometer and the soil
water. For the vadose zone, the pressure head is inherently
negative, i.e., the free water surface in the open arm of the
manometer will stand below the point of termination in the
soil.
Mercury often is used in the manometer, reducing ma-
nometer size (Figure 4-5). Other measuring devices include
vacuum gauges and pressure transducers. In areas subject to
freezing, a 40 percent ethylene-glycol solution can be used in
the tensiometer in place of water (Stephens and Knowlton,
1986).
The effective pressure range of a standard tensiometer, 0
to about -0.08 megapascals (MPa), is limited by the fact that
negative pressures are measured with reference to atmo
spheric pressure. Peck and Rabbidge (1966; 1969) developed
an osmotic tensiometer for field use that expands the effective
measurement range from O to as low as -1.5 MPa. Another
instrument that has a wide range of pressure measurements is
the thermocouple psychrometer (Table 4-5).
Hydraulic head can vary temporally at any given well.
The variation may be the result of an aquifer's response to a
known stress (e.g., a pumping well or seasonal changes in
recharge) and may demonstrate a temporal relationship be-
tween hydraulic head and contamination concentrations. For
example, an observation well, located adjacent to a ditch that
only contains water during the growing season, exhibits
changes in hydraulic head that cause seasonal changes in
uranium concentrations (Figure 4-6). This change highlights
the importance of a sampling frequency sufficient to monitor
Vacuum
Gauge
Porous
Ceramic
Cup
Figure 4-5. Schematic illustration of the essential parts of a tensiometer (from Hillel, 1980).
47
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5415 -
5410 -
-------
Table 4-6. Summary of Methods to Measure Storage Properties
Method
Application
Reference
Pumping test Can be used to measure storage values for unconfined or confined aquifers.
Multiple-well tests are more accurate than single-well tests.
Tests a relatively large volume of the aquifer.
Slug test Single-well tests for confined or unconfined aquifers. Test highly influenced by well
construction and borehole conditions.
Bureau of Reclamation
(1985); Stallman (1971);
Driscoll (1986);
Lehman (1972)
Hvorslev (1951); Bouwer and
Rice (1976); Bouwer (1989):
Lehman (1972);
Cooper et al. (1967)
Nwankwor et al. (1984);
Neuman (1972)
Nwankwor et al. (1984)
Water-balance Measures specific yield only. Requires several observation wells around pumping
well to accurately determine the cone of depression. Tests a relatively large volume
of the aquifer.
Laboratory Obtain a maximum long-term value. Fractures, macropores, and heterogeneities
of geologic material may not be represented. Only specific yield can be determined.
From Thompson et al., 1989
Copyright 1989 Electric Power Research Institute. EPRI EN-6637. Techniques to Develop Data for Hydrogeochemical Models. Reprinted with
permission.
Table 4-7. Summary of Methods to Measure Saturated Hydraulic-Conductivity Values in the Field and Laboratory (modified from
Thompson et al,
Method
Application
Reference
Slug test Confined aquifers with fully penetrating wells screened along the entire aquifer
thickness. Single-well test for wells.
Pumping test Complex multiple-well tests for confined or unconfined aquifers with fully or partially
penetrating wells. Used for wide range of aquifer permeabilities. Test wells can be
used for sampling. Tests a relatively large volume of the aquifer.
Hvorslev (1951); Bouwer and
Rice (1976); Lehman (1972)
Bureau of Reclamation
(1985); Stallman (1971);
Driscoll (1986):
Lehman (1972);
Steady-state Laboratory method to determine sample hydraulic conductivity within a range from Klute and Dirksen (1986)
permaemeter 1.0 cm/sec to 105crn/sec.
Falling-head Laboratory method to determine sample hydraulic conductivity within a range from Klute and Dirksen (1986)
permeameter 103cm/sec to 109 cm/sec.
Copyright® 1989 Electric Power Research Institute. EPRI EN-6637. Techniques to Develop Data for Hydrogeochemical Models. Reprinted with
Drawdown is defined as the drop in water level from
static-water level conditions as a result of pumping stress.
Time-drawdown and distance-drawdown data are analyzed
with model equations and type-curve matching, straight-line
matching, or inflection-point selection techniques. For ex-
amples, see Bentall (1963); Ferns et al. (1962); Kruseman and
De Ridder (1976); Lehman (1972); Neuman (1974); Reed
(1980); Stallman (1971); Walton (1962); and Walton (1970).
One disadvantage of conducting pumping tests at hazardous
waste sites is the disposal of contaminated water. Pumping
tests are valuable, however, because a relatively large portion
of the aquifer is stressed. Therefore, the hydraulic conductiv-
ity determined from an aquifer test is more representative of
spatially averaged conditions. These type of data are required
for final design considerations of a pump-and-treat system.
The slug test method consists of causing a water-level
change within a well and measuring the rate at which the
water level in the well returns to its initial level. The water-
level change can be caused either by injecting or withdrawing
a volume of water or weighted float in the well. The rate of
recovery then can be related to the hydraulic conductivity of
the surrounding aquifer material. For further information, see
Cooper et al. (1967) or Bredehoeft and Papadopulos (1980).
As indicated in Table 4-7, a disadvantage of slug tests is
that only a small volume of aquifer material is tested. If the
49
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Table 4-8. Summary of Methods to Measure Spatial Variability of Hydrogeologic Parameters
Method Application
Reference
Piezometer slug
tests
Hydraulic
conductivity from
grain size
Surface
geophysics
Borehole
geophysics
Large-scale aquifer
tests (pumping tests)
Geological mapping
of sedimentological
fades
Continuous core
Borehole flowmeter
Localized measurement, influenced by well disturbed zone.
Efficient and easy to conduct.
Samples of aquifer material required. Empirical and poor accuracy,
especially for silt and day fractions.
Direct current resistivity, electromagnetic induction, streaming potential.
Difficult to interpret and poor accuracy.
Natural gamma, gamma-gamma density, single-point resistance, neutron.
K= (0), Accuracy?
Provides bulk parameters over relatively large region.
Problems with extrapolation-geological sections above water table and
away from site.
Split-spoon sampler, samples are disturbed. Grain size analysis,
laboratory K.
Most promising. Equipment difficult to obtain.
Hvorslev (1951); Bouwer and
Rice (1976);
Lehman (1972)
Hazen (1982): Krumbein and
Monk (1942); Masch and
Denny (1966)
Zohdy et al. (1974); Sendlein
and Yazicigal (1981);
Yazicigal and Sandlein
(1982)
Serra (1984); Wheatcratt et al.
(1986); Wyllie (1963);
Patten and Bennett (1963)
Bureau of Reclamation
(1985);
Stallman (1971);
Driscoll (1986);
Lehman (1972)
Willis (1989); Leeder (1973);
Matthews (1974);
Turnbull et al. (1950)
Wolf (1988)
Rehfeldt et al. (1988);
Hufschmied (1986);
Guthrie (1986);
Kerfoot (1964)
Copyright® 1989 Electric Power Research Institute. EPRI EN-6637. Techniques to Develop Data for Hydrogeochemical Models. Reprinted with
permission.
well has been damaged (such as from a skin effect from
drilling mud), then the test may only determine the hydraulic
conductivity of the skin (Faust and Mercer, 1984). However,
at hazardous waste sites, slug tests offer many advantages,
including (1) there are no contaminated water disposal prob-
lems when a slug rod is used to displace the water, (2) a
pressure transducer can be used to measure the pressure
response in wells so that data can be collected even in fairly
permeable material, and (3) decontamination is relatively
simple, allowing as many as a dozen wells to be slug tested in
a day. The slug test method is very inexpensive and provides
a considerable amount of data on the flow characteristics of
the subsurface.
One method to determine hydraulic conductivity that is
listed in Table 4-8 is grain size analysis. Since Hazen (1892),
a number of formulas have been proposed that relate some
measure of grain size to hydraulic conductivity (for example,
Fair and Hatch, 1933; Krumbein and Monk, 1942; Masch and
Denny, 1966; and Er-Hui, 1989). These formulas are empiri-
cal with hydraulic conductivity proportional to a function of
representative grain diameters. However, these formulas are
not very accurate, and the accuracy decreases when the samples
are predominantly silt or clay. In the early stages of a field
investigation they may be very useful. They also may be
helpful in estimating hydraulic conductivity in the vadose
zone, which can be a difficult task (see Chapter 5).
Twenty years ago, when hydrologists were mainly inter-
ested in water supply, one or two pumping tests were often
sufficient to design an adequate water supply system. With the
advent of contaminant hydrology, more information is re-
quired to understand and remediate contamination distribu-
tions in the subsurface. In general, the more detailed the
investigation, the more heterogeneous the subsurface was
observed to be. Recently, much research has focused on an
improved capability to better define spatial variability and its
impact on chemical transport. Methods used to determine
spatial variability as depicted in Table 4-8 were developed
from information in Waldrop et al. (1989). Another recent
reference on this subject is Taylor et al. (1990), in which six
borehole methods are evaluated for determining the vertical
distribution of hydraulic conductivity.
Most of this discussion has focused on hydraulic conduc-
tivity; however, many of the methods for determining hydrau-
50
-------
lie conductivity also give an estimate of storage properties
(Table 4-6). Hydraulic conductivity is needed to calculate
ground-water velocities and chemical travel times. Storage
properties are also important for the following reasons: (1)
porosity is used in chemical travel-time calculations, (2) po-
rosity is used to estimate mass in place, and (3) the storage
properties determine how rapidly the flow system will re-
spond to pumpage. This latter factor is important for pump
and-treat systems where pulsed pumping is used because the
storage properties can be used to help determine the cycle
duration of pumping.
4.3 Analysis of Data
Once ground-water data are collected, they must be ana-
lyzed and interpreted. Numerous analysis tools are available,
including graphical methods, mathematical modeling, and
geostatistical techniques. Graphical methods have been used
for years. However, with the increased use of microcomputers
and software such as geographical information systems (GIS),
database management systems (DBMS) and plotting pack-
ages, it is now easy to view data via a variety of graphical
techniques. The key is to have field data readily transferred or
directly recorded on electronic/magnetic format instead of
paper.
Mathematical models have been used extensively for
ground-water analysis since the mid-1960s. Models test hy-
pothesized conceptualizations of site conditions. They often
are enhanced by data acquisition and can test the relative
importance of some information. Knowledge of the varying
importance of data can help direct the data collection. Thus,
where appropriate, using models in unison with active field
investigations can aid in characterization efficiency. Once a
model has been properly calibrated, it can make limited
predictions about future ground-water flow, contaminant trans-
port, or the effectiveness of remedial activities. A large num-
ber of models are available and are listed in van der Heijde et
al. (1988). NRC (1990) also provides an overview of model-
ing.
During the past decade, applying geostatistical principles
(i.e., structural analysis, kriging, and conditional simulation)
to interpret ground-water data has increased. Geostatistical
techniques are used to evaluate the spatial variability of
ground-water flow parameters, particularly hydraulic head
and transmissivity. A code for performing geostatistical as-
sessments is provided in Englund and Sparks (1988). The
principles of geostatistics may be appropriate for interpolating
point data to estimate the spatial distribution of certain aspects
of ground-water quality. Kriging provides a measure of the
error of estimation, which can be mapped and used to select
locations for additional sampling points. Using this approach,
a near-optimal monitoring network can be developed for a
predetermined level of reliability.
4.4 Remedial Actions
Pumping wells are part of a ground-water flow system. In
many cases, ground-water contamination is discovered be-
cause a water-supply well has become affected. These wells
create cones of depression in the potentiometric surface that
cause water to flow toward them. If that water is carrying
contaminants, they, too, will flow toward the well. When
contamination is discovered, the immediate response is to
shut the well down. This is the correct response, but doing so
changes the ground-water flow system. The potentiometric
surface adjusts to the change in source/sink term, usually
within a few days, and chemicals begin to slowly migrate to
portions of the aquifer that perhaps were previously uncon-
taminated. Therefore, an interim remedial action that should
be considered at such sites is well-head treatment. Such
treatment will bring the well back into production, minimiz-
ing the disruption to the water supply. It also will prevent the
further spread of contamimnts within the aquifer, which,
hopefully, will be consistent with any final remediation that is
conducted at the site.
Final remedial actions at hazardous waste sites are dis-
cussed in OTA (1984) and EPA (1988). Ground-water con-
tainment/cleanup options include physical containment (e.g.,
construction of low-permeability walls and caps/covers), in
situ treatment (e.g., bioreclamation), and hydraulic contain-
ment/cleanup (e.g., extraction wells and intercept trenches/
drains). To effect complete cleanup, a treatment train combin-
ing several methods may be formed.
When a pump-and-treat system is used for cleanup, con-
taminated ground-water or mobile nonaqueous phase liquids
(NAPLs) are captured and pumped to the surface for treat-
ment. This process requires locating the ground-water con-
taminant plume or NAPLs in three-dimensional space,
determining aquifer and chemical properties, designing a cap-
ture system, and installing extraction (and in some cases
injection) wells. Monitoring wells/piezometers, used to check
the effectiveness of the pump-and-treat system, are an integral
component of the system. Injection wells are used to enhance
the extraction system by flushing contaminants (including
some in the vadose zone) toward extraction wells or drains. A
pump-and-treat system may be used in combination with
other remedial actions, such as low-permeability walls, to
limit the amount of clean water flowing to the extraction
wells, thus reducing the volume of water to be treated. Pump
and-treat technology also can be used as a hydraulic barrier to
prevent offsite migration of contaminant plumes from land-
fills or residual NAPLs. The basic principle of a barrier well
system is to lower ground-water levels near a line of wells,
thus diverting ground-water flow toward the pumping wells.
Whether the objective of the pump-and-treat system is to
reduce concentrations of contaminants to an acceptable level
(cleanup) or to protect the subsurface from further contamina-
tion (containment), the system components are
A set of goals or objectives.
Engineered components such as wells, pumps, and a
treatment facility.
• Performance criteria and monitoring.
Termination criteria.
51
-------
Each of these components must be a part of the design
and evaluation of a pump-and-treat technology.
Pump-and-treat technology is appropriate for many
ground-water contamination problems (Ziegler, 1989). For
this technology to be effective, the physical-chemical subsur-
face system Must allow the contaminants to flow to the
extraction wells. The subsurface must have sufficient hydrau-
lic conductivity to allow fluid to flow readily and the chemi-
cals must be transportable by the fluid. These requirements
make the sure of pump-and-treat systems highly site specific.
Cases in which contaminants cannot readily flow to pumping
wells include
• Heterogeneous aquifer conditions where low-perme-
ability zones restrict contaminant flow toward ex-
tinction wells.
• Presence of chemicals that are sorbed or precipitated
on the soil and slowly desorb or dissolve back into
the ground water as chemical equilibrium changes in
response to the extraction process.
• Presence of immobile NAPMs that may contribute to
a miscible contaminant plume by prolonged dissolu-
tion (e.g., a separate phase gasoline at residual satu-
ration).
In these cases modifications to pump-and-treat technol-
ogy, such as pulsed pumping, maybe appropriate. Pump-and-
treat technology also may be used in combination (treatment
train) with other remedial alternatives, such as vacuum extrac-
tion and/or bioremediation. Under complex conditions, no
single technology is a panacea for subsurface remediation.
The main limitation of pump-and-treat technology is the
long time that may be required to achieve an acceptable level
of cleanup. The length of time results from the "tailing" effect
often observed with this remedial action. Tailing is the asymp-
totic decrease of contaminant concentration in water that is
removed in the cleanup process (Figure 4-7). Other potential
limitations include (1) a design that fails to contain the con-
taminant plume and allows continued migration of contami-
mnts either horizontally or vertically or, (2) operational failures
that allow the loss of containment. Typical operational prob-
lems stem from the failure(s) of surface equipment or electri-
cal and mechanical control systems; and chemical precipitation
causing plugging of wells, pumps, and surface plumbing.
Limitations are discussed further in Mackay and Cherry (1989).
Physical containment involves low-permeability barriers
such as slurry walls. Problems associated with slurry walls
may involve a difficulty with achieving design permeability
and underflow; such problems lead to loss of containment.
Slurry walls also may be used to prevent the movement of
clean water into an area being remediated by a pump-and-treat
system, thereby reducing the amount of water that needs
treatment. Slurry walls also reduce the amount of fresh ground
water that is contaminated in a pump-and-treat system. Drains
also can be used to create a hydraulic barrier. Factors that
must be considered in drain construction include health and
safety during construction, maintenance access, disposal of
excavated soils, and expected volume of water produced.
Generally, drains are used in shallow applications where low-
permeability material discourages the use of wells. Using
drains for deeper applications usually is not cost effective.
Other ground-water remedial actions are discussed in subse-
quent chapters.
4.5 Example-Conservation Chemical
Company Site
The Conservation Chemical Company (CCC) site is lo-
cated over an alluvial aquifer about 1,000 ft from the Missouri
River in Kansas City, Missouri (Figure 4-8). Formerly the site
was used to treat, store, and dispose of hazardous waste. As
may be seen, the Missouri River Valley is underlain by
deposits of alluvium with an average thickness of 90 to 95 ft.
The alluvial sediments contain interbedded clays, silts, sands,
and gravels. Although the composition varies locally, there
me some typical characteristics. Grain size increases with
depth, which reflects the depositional history of the Missouri
River. In many locations, the increasing grain size creates
three layers: (1) the uppermost layer is composed of silts and
clays; (2) the intermediate layer includes fine to medium
sands, and (3) the lowest layer is sands and gravels. The upper
layer is approximately 20-ft thick; the intermediate layer 40 to
60-ft thick; and the lower layer 30-ft thick. These alluvial
deposits overlie interlayered shales and limestones.
The alluvial aquifer is highly productive and supplies
about 500,000 gpd to a well located less than 2,000 ft from the
site. The aquifer is generally unconfined; however, short-term
responses to pumping tests and river-level variation indicate
semiconfined conditions. Various hydraulic tests conducted
on and near the site indicate that hydraulic conductivity
increases with depth, as can be expected from the grain size
distribution. Crabtree and Malone (1984) obtained hydraulic
conductivity estimates from 0.51 to 2.35 ft/d for the shallow
alluvium. Pumping tests at a nearby production well indicated
an overall transmissivity of the aquifer between 4,000 and
16,700 ftVd and a specific yield between 0.15 to 0.27. Slug
tests were attempted but proved unsatisfactory because of
large oscillations (see Chapter 6). Analysis of the response of
the aquifer to changes in river levels suggests that the ratio of
horizontal to vertical hydraulic conductivity is about 100:1 for
the site vicinity.
Water levels are from 5 to 15 ft below land surface
(Crabtree and Malone, 1984). Water-level data indicate that
for the area south of the river, ground-water discharges to the
river; however, during periods when the river is high, ground
water flows from the river into the aquifer. This variability is
indicated in Figure 4-9 where the vector direction indicates
the flow direction and its length indicates the gradient magni-
tude. These data were collected over a 1-year period. Cluster
wells indicate a very small vertical hydraulic gradient.
The site was contaminated with metals and organic
compounds. The spatial distribution of concentrations for
specific contaminants did not define a meaningful "plume."
However, concentration of all contaminants tends to decrease
with distance from the site. Also, organic contaminants are
generally located directly under, northeast, and southeast of
52
-------
•s -s
1 1
11
o o
11
-5 i
« IS
Water Filled Aquifer Volumes
Figure 4-7. Effects of tailing on pumping time (from Keeley et al., 1989).
Sugar Creek (
Industrial Area
Mobay
^~ Chemical Co.
r— CCC Site
\KCPL
Pleasanton Group —I
Kansas City Group —I
Physiographic block diagram showing general
relationship between geologic units,
geomorphic features, site locations and other
important landmarks.
0 2000 4000 6000 8000 Feet
0 500 WOO 1500 2000 Meters
Figure 4-8. Bock diagram showing the location of the CCC site and generalized geology.
53
-------
0.004
Figure 4-9. Ground-water flow directions and gradients observed in various piezometers (from Larson, 1986).
54
-------
the site; concentrations of metals are found north and west of
the site. For the nearby offsite wells, the highest concentration
of organics generally are found in the deeper wells.
The design of a remedial pumping system at the CCC site
was complicated by two factors-the impact of the Missouri
River and the high productivity of the aquifer below the site.
Changes in river stage cause significant variations in ground-
water flow rates and directions. Consequently, the operating
system must be flexible enough to track these changes and to
modify pumping as necessary to meet design objectives.
Pumping rates required to achieve design goals are relatively
high for all the alternatives considered because of the high
productivity of the aquifer. Even with high pumping rates, the
area of influence or control is difficult to verify because the
changes in water-level elevation, normally used to determine
flow direction, are small and difficult to measure.
To evaluate optimal pumping and monitoring strategies,
an analytical approach was embedded in a linear program.
This approach accounts for variations in flow directions and
provides an analysis of pumping requirements under altern-
ative performance criteria. Hydraulic gradients are of particular
interest because performance monitoring of site pumping is
based on the measurement of water-level elevation differ-
ences between piezometer pairs. The amount of water pumped
has been minimized while performance requirements con-
tinue to be met. For a site pumping remedy, the quantifiable
performance requirement is a minimum inward hydraulic
gradient at paired piezometers.
Numerous simulations were performed using gradient
data provided in Figure 4-9. These simulations were per-
formed on both regional and local scales. The regional analy-
sis was performed to study the influence of offsite pumping
centers. Based on these simulations, a recovery system that
met all the requirements is currently being implemented.
4.6 References
Barcelona, Ml, J.P. Gibb, and R.A. Miller. 1983. A Guide to
the Selection of Materials for Monitoring Well Construc-
tion and Ground-Water Sampling. ISWS Contract Report
327. Illinois State Water Survey, Champaign, IL.
Bear, J. 1979. Hydraulics of Ground-water. McGraw-Hill,
New York.
Bentall, R. (Compiler). 1963. Shortcuts and Special Problems
in Aquifer Tests. U.S. Geological Survey Water Supply
Paper 1545-C.
Bouwer, H. 1978. Ground-Water Hydrology. McGraw-Hill,
New York.
Bouwer, H. 1989. The Bouwer and Rice Slug Test—An
Update. Ground Water 27(3):304309.
Bouwer, H. and P.C. Rice. 1976. A Slug Pest for Determining
Hydraulic Conductivity of Unconfined Aquifers with
Completely or Partially Penetrating Wells. Water Re-
sources Research 12(3):423-428.
Bredehoeft, J.D. and S.S. Papadopulos. 1980. A Method for
Determining the Hydraulic Properties of Tight Forma-
tions. Water Resources Research 16(l):233-238.
Bureau of Reclamation. 1985. Ground-Water Manual-A
Water Resources Technical Publication, 2nd ed. U.S.
Department of the Interior, Bureau of Reclamation, Den-
ver, CO.
Campbell, G.S. and G.W. Gee. 1986. Water Potential: Miscel-
laneous Methods. In: Methods of Soil Analysis, Part 1,
2nd ed., A. Klute (ed.), Agronomy Monograph No. 9,
American Society of Agronomy, Madison, WI, pp. 619-
633.
Cassel, O.K. and A. Klute. 1986. Water Potential: Tensiom-
etry. In: Methods of Soil Analysis, Part 1, 2nd ed., A.
Klute (ed.), Agronomy Monograph No. 9, American So-
ciety of Agronomy, Madison, WI, pp. 563-596.
Cooper, H. H., Jr., J.D. Bredehoeft, and I.S. Papadopulos.
1967. Response of a Finite-Diameter Well to an Instanta-
neous Charge of Water. Water Resources Research
3(l):263-269
Crabtree, J.E. and P.G. Malone. 1984. Hydrogeologic Charac-
terization Conservation Chemical Company Site, Kansas
City, Missouri. U.S. Army Corps of Engineers Water-
ways Experiment Station, Vicksburg, MS.
Davis, S .N. and R.J. DeWiest. 1966. Hydrogeology. John
Wiley & Sons, New York, 463 pp.
DeWiest, R.J.M. 1969. Flow through Porous Media. Aca-
demic Press, New York.
Domenico, P. A. 1972. Concepts and Models in Ground-water
Hydrology. McGraw-Hill, New York.
Driscoll, F.G. 1986. Ground-water and Wells, 2nd ed. John-
son Division, St. Paul, MN, 1089 pp.
Englund, E. and A. Sparks. 1988. GEO-EAS (Geostatistical
Environmental Assessment Software) User's Guide. EPA/
600/4-88/033a (Guide: NTIS PB89-151252, Software:
NTIS PB89-151245).
Er-Hui, Z. 1989. Experimental Research on Permeability of
Granular Media. Ground-Water, 27(6):848-854.
Fair, G.M. and L.P. Hatch. 1933. Fundamental Factors Gov-
erning the Streamline Flow of Water through Sand. J.
Am. Water Works Ass. 25:1551-1565.
Faust, C.R. and J.W. Mercer. 1984. Evaluation of Slug Tests
in Wells Containing a Finite-Thickness Skin. Water Re-
sources Research 20(4):504-506.
Ferris, J.G., D.B. Knowles, R.H. Brown, and R.W. Stallman.
1962. Theory of Aquifer Tests. U.S. Geological Survey
Water Supply Paper 1536-E, 174 pp.
55
-------
Fetter, Jr., C.W. 1981. Determination of the Direction of
Ground-Water Flow. Ground-Water Monitoring Review
1 (3):28-31.
Freeze, R.A. and J.A. Cherry. 1979. Ground-water. Prentice-
Hall, Englewood Cliffs, NJ.
Garber, M.S. and F.C. Koopman. 1968. Methods of Measur-
ing Water Levels in Deep Wells. U.S. Geological Survey
Techniques of Water-Resources Investigations TWI 8-
Al.
GeoTrans, 1989. Ground-water Monitoring Manual for the
Electric Utility Industry. Edison Electric Institute, Wash-
ington, DC.
Goode, D.J. andR.J. Wilder. 1987. Ground-water Contamina-
tion Near a Uranium Tailing Disposal Site in Colorado.
Ground-Water 25(5):545-554
Guthrie, M. 1986. Use of a Geoflowmeter for the Determina-
tion of Ground-Water Flow Direction. Ground-Water
Monitoring Review 6(1):81.
Hazen, A. 1892. Experiments upon the Purification of Sewage
and Water at the Lawrence Experiment Station. In: 23rd
Annual Report, Massachusetts State Board of Health.
Hillel, D. 1980. Fundamentals of Soil Physics. Academic
Press, New York.
Hufschmied, P. 1986. Estimation of Three-Dimensional Sta-
tistically Anisotropic Hydraulic Conductivity Field by
Means of Single Well Pumping Tests Combined with
Flowmeter Measurements. Hydrogeologic 1986(2): 163-
174.
Hvorslev, M.J. 1951. Time Lag and Soil Permeability in
Ground-water Observations. U.S. Army Corps of Engi-
neers Waterways Experiment Station, Bull. 36, Vicksburg,
MS.
Keeley, J.W., D.C. Bouchard, M.R. Scalf, and C.G. Enfield.
1989. Practical Limits to Pump and Treat Technology for
Aquifer Remediation. Submitted to Ground-Water Moni-
toring Review.
Kerfoot, W.B. 1984. Darcian Flow Characteristics Upgradient
of a Kettle Pond Determined by Direct Ground-Water
Flow Measurement. Ground-Water Monitoring Review
4(4):188-192.
Klute, A. and C. Dirksen. 1986. Hydraulic Conductivity and
Diffusivity: Laboratory Methods. In: Methods of Soil
Analysis, Part 1, 2nd ed., A. Klute (ed.), Agronomy
Monograph No. 9, American Society of Agronomy, Madi-
son, WI, pp. 687-734.
Krumbein, W.C. and G.D. Monk. 1942. Permeability as a
Function of the Size Parameters of Unconsolidated Sand.
AIMME Petroleum Division Technical Pub. No. 1492
(published in Petroleum Technology Vol. 5, No. 4).
Kruseman, G.P. andN.A. De Ridder. 1976. Analysis and
Evaluation of Pumping Test Data. International Institute
for Land Reclamation and Improvement, Wageningen,
The Netherlands, 200 pp.
Larson, D. 1981. Materials Selection for Ground-Water Moni-
toring. Paper Presented at the National Water Well Asso-
ciation Short Course entitled Practical Considerations in
the Design and Installation of Monitoring Wells, Colum-
bus, OH, December 16-17.
Larson, S.P. 1986. Notes from November Meeting between
Settling Defendants and EPA, Washington, DC.
Leeder, M.R 1973. Fluviatile Fining-Upward Cycles and the
Magnitude of Paleochannels. Geology Magazine
110(3):265-276.
Lohman, S.W. 1972. Ground-water Hydraulics. U.S. Geo-
logical Survey Professional Paper 708.
Mackay, D.M. and J.A. Cherry. 1989. Ground-Water Con-
tamination: Pump-and-Treat Remediation. Environ. Sci.
Technol. 23(6):630-636.
Masch, F. and K. Denny. 1966. Grain Size Distribution and Its
Effect on the Permeability of Unconsolidated Sands.
Water Resources Research 2(4):665-577.
Matthews, R.K. 1974. Dynamic Stratigraphy. Prentice Hall,
Englewood Cliffs, NJ, Chapter 10, pp. 137-172.
National Research Council (NRQ. 1990. Ground-Water Mod-
els Scientific and Regulatory Applications. National Acad-
emy Press, Washington, DC, 303 pp.
Neuman, S.P. 1972. Theory of Flow in Unconfined Aquifers
Considering Delayed Response of the Water Table. Wa-
ter Resources Research 8(4): 1031-1045.
Neuman, S.P. 1974. Effect of Partial Penetration on Flow in
Unconfined Aquifers Considering Delayed Gravity Re-
sponse. Water Resources Research 10(2):303-312.
Nwankwor, G. I., J.A. Cherry, and R.W. Gillman. 1984. A
Comparative Study of Specific Yield Determinations for
a Shallow Sand Aquifer. Ground-Water 22(6):764-772.
Office of Technology Assessment (OTA). 1984. Protecting
the Nation's Ground-water from Contamination. OTA-0-
233. U.S. Office of Technology Assessment, Washing-
ton, DC.
Patten, Jr., E.P. and G.D. Bennett. 1963. Application of Elec-
trical and Radioactive Well Logging to Ground-Water
Hydrology. U.S. Geological Survey Water Supply Paper
1544-D, 60 pp.
Peck, A.J. and R.M. Rabbidge. 1966. Soil Water Potential:
Direct Measurements by a New Technique. Science
151:1385-1386.
56
-------
Peck, AJ. and R.M. Rabbidge. 1969. Design and Perfor-
mance of an Osmotic Tensiometer for Measuring Capil-
lary Potential. Soil Sci. Soc. Am. Proc. 33:196-202.
Phene, CJ. and D.W. Beale. 1976. High-Frequency Irrigation
for Water-Nutrient Management in Humid Regions. Soil
Sci. Soc. Am. J. 40:430-436.
Rawlins, S.L. and G.S. Campbell. 1986. Water Potential:
Thermocouple Psychrometry. In: Methods of Soil Amly-
sis, Part 1, 2nd ed., A. Klute (ed.), Agronomy Monograph
No. 9, American Society of Agronomy, Madison, WI, pp.
597-618.
Reed, I.E. 1980. Type Curves for Selected Problems of Flow
to Wells in Confined Aquifers. U.S. Geological Survey
Techniques of Water-Resources Investigations TWI 3-
B3.
Rehfeldt, K.R, P. Hufschmied, L.W. Gelhar, and M.E.
Schaefer. 1988. The Borehole Flowmeter Technique for
Measuring Hydraulic Conductivity Variability. Draft topi-
cal report prepared by MIT for Electric Power Research
Institute, Research Project 2485-5.
Rehm, B. W., B.J. Christel, T.R. Stolzenburg, D.G. Nichols,
B. Lowery, and B.J. Andraki. 1987. Field Evaluation of
Instruments for the Measurement of Unsaturated Hydrau-
lic Properties of Fly Ash. EPRI EA-5011. Electric Power
Research Institute, Palo Alto, CA.
Scalf, M.R., J.F. McNabb, W.J. Dunlap, R.L. Cosby, and J.
Fryberger. 1981. Manual of Ground-Water Quality Sam-
pling Procedures. EPA/600/2-81/160, (NTIS PB82-
103045). Also published in NWWA/EPA Series, National
Water Well Association, Dublin OH.
Sendlein, L.V.A. and H. Yazicigal. 1981. Surface Geophysi-
cal Techniques in Ground-Water Monitoring, Part 1.
Ground-Water Monitoring Review 1(4): 42-46.
Serra, 0. 1984. Fundamentals of Well-Log Interpretation, 1:
The Acquisition of Logging Data. In: Developments in
Petroleum Science, Vol. 15A. Elsevier, New York, 423
PP.
Sisk, S.W. 1981. NEIC Manual for Ground-water/Subsurface
Investigations at Hazardous Waste Sites. EPA/330/9-81-
002 (NTIS PB82-103755).
Stallman, R.W. 1971. Aquifer-Test Design, Observation, and
Data Analysis. U.S. Geological Survey Techniques of
Water-Resources Investigations TWI 8-B1.
Stannard, D.I. 1986. Theory, Construction and Operation of
Simple Tensiometers. Ground-Water Monitoring Review
6(3):70-78.
Stephens, D.B. and R. Knowlton, Jr. 1986. Soil Water Move-
ment and Recharge through Sand at a Semiarid Site in
New Mexico. Water Resources Research 22(6):881-889.
Taylor, K., S. Wheatcraft, J. Hess, J. Hayworth, and F. Molz.
1990. Evaluation of Methods for Determining the Verti-
cal Distribution of Hydraulic Conductivity. Ground-Wa-
ter 28(l):88-98.
Thompson, C.M., et al. 1989. Techniques to Develop Data for
Hydrogeochemical Models. EPRI EN-6637. Electric
Power Research Institute, Palo Alto, CA.
Thomhill, J.T. 1989. Accuracy of Depth to Water Measure-
ments. EPA Superfund Ground-Water Issue Paper. EPA/
540/4-89/002.
Todd, O.K. 1980. Ground-water Hydrology. John Wiley &
Sons, New York.
Tumbull, W.J., E.S. Krinitsky, and L.J. Johnson. 1950. Sedi-
mentary Geology of the Alluvial Valley of the Missis-
sippi River and its Bearing on Foundation Problems. In:
Applied Sedimentation, P.O. Trask, (ed.), John Wiley &
Sons, New York, pp. 210-226.
U.S. Environmental Protection Agency (EPA). 1986. RCRA
Ground-Water Monitoring Technical Enforcement Guid-
ance Document. EPA OSWER-9950. 1. Also published in
NWWA/EPA Series, National Water Well Association,
Dublin, OH.
U.S. Environmental Protection Agency (EPA). 1987. Hand-
book Ground-water. EPA/625/6-87/016.
U.S. Environmental Protection Agency (EPA). 1988. Guid-
ance on Remedial Actions for Contaminated Ground-
Water at Superfund Sites. Advance Copy, OSWER
Directive No. 9283.1-2.
U.S. Geological Survey (USGS). 1977. National Handbook of
Recommended Methods for Water Data Acquisition
(Chapter 2-Ground-Water, updated January 1980).
USGS Office of Water Data Coordination, Reston, VA.
van der Heijde, P.K M., A.I. El-Kadi, and S.A. Williams.
1988. Ground-water Modeling An Overview and Status
Report. EPA/600/2-89/028.
Waldrop, W.R., K.R. Rehfeldt, L.W. Gelhar, J.B. Southard,
and A.M. Dasinger. 1989. Estimates of Macrodispersivity
Based on Analyses of Hydraulic Conductivity Variability
at the MADE Site. EPRI EN-6405. Electric Power Re-
search Institute, Palo Alto, CA.
Walton, W.C. 1962. Selected Analytical Methods for Well
and Aquifer Evaluation. ISWS Bulletin 49. Illinois State
Water Survey, Champaign, IL.
Walton, W.C. 1970. Ground-water Resource Evaluation.
McGraw-Hill, New York.
Walton, W.C. 1987. Ground-water Pumping Tests Design and
Analysis. Lewis Publishers, Chelsea, MI, 201 pp.
57
-------
Wheatcraft, S.W., K.C. Taylor, J.W. Hess, and T.M. Morris. Yazicigal, H. and L.V.A. Sendlem. 1982. Surface Geophysi-
1986. Borehole Sensing Methods for Ground-Water In- cal Techniques in Ground-Water Monitoring, Part II.
vestigations at Hazardous Waste Sites. EPA/600/2-86/11 Ground-Water Monitoring Review 2(l):56-62.
(NTIS PB87-132783).
Ziegler, G.J. 1989. Remediation through Ground-Water Re-
Willis, B J. 1989. Paleochannel Reconstruction from Pointbar covery and Treatment. Pollution Engineering 22(7):75-
Deposits: A Three-Dimensional Perspective. Sedimen- 79.
tology 36:757-766.
Zohdy, A.A.R., G.P. Eaton, and D.R. Mabey. 1974. Applica-
Wolf, S. 1988. Spatial Variability of Hydraulic Conductivity tion of Surface Geophysics to Ground-Water Investiga-
in a Sand and Gravel Aquifer. Master of Science Thesis, tions. U.S. Geological Survey Techniques of
Department of Civil Engineering, MIT, Cambridge, MA. Water-Resources Investigations TWI 2-DI.
Wyllie, M.R.J. 1963. The Fundamentals of Well Log Interpre-
tation. Academic Press, New York, 238 pp.
58
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Chapter 5
Characterization of the Vadose Zone
James W. Mercer and Charles P. Spalding
The vadose zone is the subsurface extending from land
surface to the water table. It also is called the zone of aeration,
variably saturated zone, or the unsaturated zone. Use of this
latter term is discouraged, however, because since the vadose
zone contains moisture up to 100 percent saturation, the term
unsaturated could be misleading. Depth of the vadose zone
can vary greatly depending on the region of the site. For
example, in the humid eastern portion of the United States, the
vadose zone can be only a few feet thick, disappearing during
times of the year when the water table is high. In the arid west,
the vadose zone can be several hundred feet thick.
Because the vadose zone overlies the saturated zone,
chemical releases at or near the land surface must pass through
the vadose zone before reaching the water table. Therefore, at
many contaminated sites, often both the vadose zone and the
saturated zone need to be characterized and remediated (i.e.,
treatment trains must be applied). As discussed later in this
chapter, the vadose zone can have more complex flow condi-
tions than the saturated zone. These conditions can be difficult
to characterize. On the other hand, because the vadose zone is
nearer to the land surface, for remedial actions, the flow
system may not need to be completely characterized under
certain site conditions and contaminants.
The main difference between the saturated and vadose
zones is the presence of air/gas in the pore spaces of the
vadose zone. The amount of water and air varies both spatially
and temporally, which contributes to the complex nature of
the vadose-zone flow system. However, the presence of soil
gas also provides a valuable screening tool for locating vola-
tile organic compounds (VOCs) In addition, there is the
potential for significant biological activity. The advantages
and disadvantages of characterizing and remediating the va-
dose zone are discussed in the following sections: (1) Review
of Concepts, (2) Field Techniques, (3) Analysis of Data, and
(4) Remedial Actions. An example of the application of
techniques as discussed in the chapter follows these sections.
5.1 Review of Concepts
The vadose zone can be divided into (1) the belt of soil
water, (2) the intermediate belt, and (3) the capillary fringe.
The belt of soil water is the uppermost zone extending from
the land surface to a depth where soil moisture changes are
minimal. It contains the root zone of plants, and is the site of
many active processes. Precipitation, for example, falls to the
land surface and runs off via overland flow or infiltrates into
the ground. Working against the infiltrating water are evapo-
ration and transpiration. Evaporation is the process that con-
verts the water at or near land surface to vapor. Transpiration
is the process by which plant roots absorb water and release
water vapor back to the atmosphere through their leaves and
stems. Hydrologists combine these two processes into the
term evapotranspiration. Much of the infiltrating water is
consumed by evapotranspiration. The water that is not con-
sumed and eventually makes it to the water table is recharge.
It is important to understand and characterize these processes
for hazardous waste sites (1) to help understand recharge
events and how contaminants may move through the vadose
zone, and (2) to help design caps used to limit infiltration and
recharge to a contaminant source area.
The capillary fringe (at the base of the vadose zone)
extends upward from the water table until there is a decrease
in soil moisture. Portions of this zone can be at 100 percent
saturation. This zone also will change as recharge/discharge
causes the water table to fluctuate. The capillary fringe is
formed due to a capillary rise caused by the surface tension
between air and water. Hydraulic head is made up of an
elevation head and a pressure head. At the water table, the
pressure head is zero. It increases below the water table and
decreases above the water table. That is, pressure head is
negative in the vadose zone, a phenomenon sometimes re-
ferred to as soil tension or suction. The latter term refers to the
effect of water being sucked into a dry soil. The negative
pressure head will pull water upward from the saturated zone,
forming the capillary fringe. The height of the capillary fringe
depends on the pore size of the soil (e.g., the capillary rise is
greater for smaller pores). Unfortunately, pore size is difficult
to determine and is not directly related to grain size.
Hydraulic head in the vadose zone is defined the same
way as it is in the saturated zone-the sum of pressure head
and elevation head. In the vadose zone, however, pressure
head is used for the saturation-dependent relationships. Capil-
lary pressure, defined as the difference between the nonwetting
fluid pressure and the wetting fluid pressure, also is used. For
an air-water system, the air pressure is assumed to be negli-
gible, and capillary pressure is essentially equal to the nega-
tive of the pressure head.
The moisture present in the vadose zone is quantified by
a term called the volumetric water content or degree of
59
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Legend
Clay
Clay Loam
Loam
Loamy Sand
Silt Loam
Silty Clay Loam
Sand
Sandy Clay
Sandy Clay Loam
Sandy Loam
X----X
*----*
B----H
«----*
10
0.2 0.3
Volumetric Water Content
Figure 5-1. Moisture characteristic or specific retention curves for various soil types.
60
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saturation. Saturation varies from zero to one and refers to the
amount of volume of pore space filled with water. Volumetric
water content varies between zero and the porosity value. For
complete saturation, the volumetric water content is equal to
porosity, and the degree of saturation is 100 percent or 1.0. If
the pore space is only half filled with water, then the satura-
tion is 50 percent or 0.5 and volumetric water content is half
the porosity.
In the vadose zone, a relationship called the moisture
characteristic curve exists between volumetric water content
and pressure head (Figure 5-1). As the figure shows, this
curve is nonlinear and generally is not a single-valued-func-
tion relationship. That is, a different curve is used to describe
the pressure-head-volumetric-water-content relationship de-
pending on whether the soil is filling or draining. Depending
on the wetting history, an entire set of curves is needed. This
phenomenon is called hysteresis, and is due in part to en-
trapped air in the soil after wetting. This set of curves is
necessary to fully describe the flow conditions in the vadose
zone.
Flow in the vadose zone is complicated further by the
presence of air. Because both air and water are in the pore
space, each resists the flow of the other. This results in a
decrease in fluid mobility, characterized by the term relative
permeability. Relative permeability varies between zero and
one. It is a nonlinear function of saturation that also can
exhibit hysteresis. Thus, to fully characterize flow in the
vadose zone, the relative permeability function must be known,
in addition to the saturated hydraulic conductivity.
5.2 Field Techniques
Based on the review of concepts, near-surface processes,
as well as other parameters that are functions of moisture
content, need to be characterized. For hazardous waste reme-
diation, vadose zone processes must be understood to design
caps and covers to minimize infiltration. Methods to measure
or estimate these processes/parameters are discussed in this
section. Reviews of vadose zone monitoring are discussed in
Wilson (1980, 1981, 1982, 1983). Section 9.2 further dis-
cusses sampling of subsurface solids and vadose zone water,
and Table 9-5-identifies additional references focusing on
characterization of the vadose zone.
5.2.1 Precipitation and Infiltration
Precipitation is defined as the total amount of water that
reaches land surface, and is measured with gauges as a depth
of water (see Table 5-1). Because weather stations are not
generally set up at hazardous waste sites, precipitation infor-
mation is obtained from nearby airports. Another source of
precipitation data is the National Climatic Data Center in
Asheville, North Carolina. Wind velocity and air temperature
also are studied for remediation.
The maximum rate at which water can enter a soil is the
infiltration capacity or potential infiltration rate. The maxi-
mum rate occurs when the water supply at the surface is
unlimited. During precipitation events, all the water will
infiltrate if the rainfall intensity is less than the infiltration
capacity. If this capacity is exceeded, the excess rain cannot
infiltrate and will produce surface runoff. Although this dis-
cussion concerns water infiltration, it also applies to a chemi-
cal spill infiltrating the subsurface. Infiltration capacity varies
with time; it is highest at the begiming of a precipitation event
and decreases as the soil becomes saturated. Table 5-2 lists
methods to measure or estimate infiltration rates. These meth-
ods are discussed in Thompson et al. (1989) and in the
references provided in the table.
Spatial variability is present in the vadose zone as well as
the saturated zone. Spatial variability produces a fingering of
flow as it moves downward from the surface. This means that
the wetting front does not move as a sharp front, but instead
moves downward with an irregular shape where some zones
(fingers) move more rapidly than other zones. Laboratory
studies by Stephens and Heermann (1988) suggest that this
variability increases with decreasing soil moisture content.
5.2.2 Evaporation and Evapotranspiration
Evaporation is the loss of water from the soil into the
atmosphere. In the absence of vegetative cover, the bare soil
surface is subject to radiation and wind effects, and soil water
evaporates directly from the soil surface. An associated pro-
cess is evaporation of water from plants, or transpiration. For
evaporation to occur (1) a continual supply of heat must meet
the latent heat requirements, (2) a vapor pressure gradient
must exist between the soil surface and the atmosphere, and
(3) there must be a continual supply of water from and/or
through the soil layers. The first two conditions determine the
evaporative demand (Table 5-3) and are controlled by micro-
meteorological factors such as air temperature, humidity,
Table 5-1. Summary of Methods to Measure Precipitation
Method Application
Reference
Sacramento gage
Weighing gage
Tipping-bucket gage
Accumulated precipitation. Manual recording.
Continuous measurement on precipitation. Mechanical recording.
Continuous measurement of precipitation. Electronic recording.
Recommended.
Finkelstein et al. (1989);
National Weather Service (972)
Finkelstein et al. (1989)
Finkelstein et al. (1989)
From Thompson et al., 1989
Copyright® 1989 Electric Power Research Institute. EPRI EN-6637. Techniques to Develop Data for Hydrogeochemical Models. Reprinted with
permission.
61
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Table 5-2 Summary of Methods to Measure or Estimate infiltration Rates
Method Application
Reference
Infiltrometers Measures the maximum infiltration rate of surface soils. Useful for determining
relative infiltration rates of different soil types: however, infiltration rates
determined by this method tend to overestimate actual rates.
Sprinkler Measures the potential range of infiltration rates under various precipitation
infiltrometer conditions. Tends to be expensive and non-portable. Sprinkler infiltrometers
have typically been used for long duration research studies.
Dunne and Leopold (1978);
Bouwer (1986)
Dunne and Leopold (1978);
Peterson and Bubenzer
(1986)
Average infiltration
method
Empirical
relations
infiltration
equations
Method for estimating the average infiltration rate for small watersheds.
Provides an approximate estimate of infiltration for specific precipitation events
and antecedent moisture conditions.
Methods to approximate the infiltration for large watersheds. These methods can be
useful when combined with limited infiltrometer measurements to obtain a gross
approximation of infiltration.
Analytical equations for calculating infiltration rates. Parameters required in the
equations can be readily measured in the field or obtained from the literature.
Probably the least expensive and most efficient method for estimating infiltration.
Dunne and Leopold (1978)
Musgrave and Holtan (1964)
Bouwer (1986);
Green and Ampt (1911);
Philip (1957)
From Thompson et a/., 7989
Copyright® 1969 Electric Power Research Institute. EPRI EN-6637. Techniques to Develop Data for Hydrogeochemical Models. Reprinted with
permission.
Table 5-3. Summary of Methods to Measure Evaporation
Method Application
Reference
Class-A pan
Weighing lysimeter
Remote sensing
Evaporation from surface of free liquid.
Direct measure of bare soil evaporation.
Currently in development. Useful for large areas.
Veihmeyer (1964);
National Weather Service (1972)
USGS (1977) (updated 1982)
USGS (1977) (updated 1982)
Modified from Thompson et a/., 7989
Copyright® 1989 Electric Power Research Institute. EPRI EN-6637. Techniques to Develop Data for Hydrogeochemical Models. Reprinted with
wind velocity, radiation, and crop cover. The third condition,
which determines the rate of water supply to the evaporative
site (soil surface), is controlled by soil-water content, pressure
potential, and relative permeability of the soil. Thus, the
actual evaporation rate is determined by evaporative demand
and soil hydraulic properties.
Transpiration occurs in response to a vapor pressure
deficit between leaves and the atmosphere. To meet this
demand, plants must extinct water from the root zone. Com-
bined losses due to evaporation and transpiration are com-
monly referred to as evapotranspimtion. When the soil surface
is covered completely by a crop canopy, evaporation losses
are negligible, and transpiration is the principal process by
which water is lost from the root zone. The same environmen-
tal factors that control evaporation also control the potential
transpiration. Table 5-4 summarizes methods to measure or
estimate evapotranspiration.
5.2.3 Moisture Content and Moisture
Characteristic Curves
In the vadose zone, the void space is partly filled by air
and partly by water. The moisture content or volumetric water
content represents the quantity of water present at a certain
time at a point in the porous media. The maximum value of
volumetric water content occurs when all voids are filled; the
minimum value occurs when all voids are empty (filled with
air). Thus, moisture content varies between 0 to the value of
the soil porosity.
Changes in moisture content are important to detect. For
example, under a cap/cover, changes in moisture content
could indicate leaks in the cover. By determining moisture
content with depth, perched water zones can be located for use
in water quality sampling. Several methods are used to mea-
sure moisture content (see Table 5-5), but the recommended
techniques are gravimetric and neutron scattering. Gravimet-
ric moisture content measurements are made by weighing
soils before and after drying. The neutron scatter method
lowers the moisture meter, which contains a source of fast
neutrons and a slow neutron detector, into the soil through an
access tube (Figure 5-2). Neutrons are emitted by the source
(e.g., radium or americium-beryllium) at a very high speed.
When these neutrons collide with a small atom, such as
hydrogen contained in soil water, their direction of movement
is changed and they lose part of their energy. These "slowed"
neutrons are measured by a detector tube and a scalar. This
62
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Table 5-4. Summary of Methods to Measure or Estimate Evapotranspiration
Method Application
Reference
WATER BALANCE METHODS
Pan lysimeter
Soil moisture sampling
Direct field method; accurate; moderate to low cost.
Direct field method; accurate; moderate to low cost.
Potential evapotranspirometers Direct field method of PET Moderately accurate and
low cost.
Cl tracer
Water-budget analysis
Ground-water fluctuation
Indirect combined field and laboratory method;
moderate to high cost.
Indirect field estimate of ET; manageable to difficult;
moderate to low cost.
Indirect field method; moderate to low cost.
Veihmeyer (1964);
Sharma (1985)
Veihmeyer (1984)
Thornthwaite and Mather (1955)
Sharma (1985)
Davis and DeWiest (1966)
Davis and DeWiest (1966)
MICROMETEOROLOGICAL METHODS
Profile method Indirect field method.
Energy budget/
Bowen ratio
Eddy covariance method
Penman equation
Thornthwaite equation
Blaney-Criddle equation
Indirect field method; difficult; costly; requires data
which is often unobtainable; research oriented.
Indirect field method; costly measures water-vapor fiux
directly; highly accurate; well accepted; research oriented.
Indirect field method; difficult; costly; very accurate;
eliminates need for surface temperature measurements;
research oriented.
Empirical equation; most accepted for calculating PET:
uses average monthly sunlight: moderate to low cost.
Empirical equation; widely used; moderate to high
accuracy; low cost; adjusts for certain crops and vegetation.
Sharma (1985)
Veihmeyer (1964);
Shamra (1985)
Veihmeyer (1964); Sharma (1985)
Veihmeyer (1964); Sharma (1985)
Veihmeyer (1964); Sharma (1985)
Stephens and Stewart (1964)
From Thompson et a/., 7989
Copyright® 1989 Electric Power Research Institute. EPRI EN-6637. Techniques to Develop Data for Hydrogeochemical Models. Reprinted with
permission.
Table 5-5. Summary of Methods for Measuring Moisture Content
Method Application
Reference
Gravimetric
Neutron scattering
Gamma ray
attenuation
Electromagnetic
Tensiometry
Laboratory measurements of soils which should be dried at
110°C. The standard method for moisture content determination.
Recommended.
In situ measurements via installed access tubes. Widely used.
Requires calibration curves. Recommened.
In situ measurements via installed access tubes. Difficult to use.
Not recommended for routine use.
In situ measurements from implanted sensors. Not widely used.
Not recommended for routine use.
Gardner (1986):
Radian Corporation (1988)
van Save/ (1963)
Gardner (1986)
Schmugge et a/. (1980)
In situ measurements inferred from moisture-matric potential relationship. Gardner (1986)
Prone to error resulting from uncertainty of moisture-matric potential
relationship. Not recommended.
From Thompson et a/., 7989
Copyright® 1989 Electric Power Research Institute. EPRI EN-6637. Techniques to Develop Data for Hydrogeochemical Models. Reprinted with
permission.
63
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Cable
Sealer
Shield and
Standard
Surface
f *
Access Tube -?—*
«-.•"'..' C-5 ' •£
Figure 5-2.
Components of the neutron moisture meter (from
Millet, 1980).
reading is then related to the soil moisture content in the
vadose zone and porosity in the saturated zone. These mea-
surements are good indicators of relative changes in moisture
content; absolute values of moisture content are difficult to
determine.
If the moisture characteristic curve is known (Figure 5-2),
then pressure head can be measured using, for example, a
tensiometer, and then converted to moisture content using the
characteristic curve. Because of the uncertainty involved,
however, this approach is not recommended.
In a saturated soil at equilibrium with free water at the
same elevation, the matric potential or negative pressure
potential is atmospheric and hence equal to zero.
Subatmospheric pressure (suction or tension) applied to soil
draws water out of the soil, as the voids cannot retain water
against the applied suction. Thus, increasing matric potential
is associated with decreasing volumetric water content. The
soil-water retention curve, also known as the soil-water char-
acteristic curve, expresses this relationship. In most soils,
drainage (drying) and infiltration (wetting) produce different
water retention curves (Figure 5-3). This is because air that is
trapped in the pores upon wetting decreases the water content.
In this case, the soil-water characteristic curve is said to
display hysteresis. Table 5-6 lists methods for determining
moisture characteristic curves.
5.2.4 Vadose-Zone Hydraulic Conductivity
The hydraulic conductivity of a porous medium is largest
at saturation and decreases as the water content decreases. The
saturated hydraulic conductivity in the vadose zone, as well as
the relationship between water content and hydraulic conduc-
tivity, must be determined. At relatively low water contents,
the hydraulic conductivity decreases primarily because air
occupies more of the pore space, leaving less cross-sectional
area available for water transport. The film of water covering
the soil particles becomes thinner and thinner, until at low
water contents, it becomes thin enough that attractive forces
between the water molecules and the soil particles become
stronger than other forces that might be acting to make water
move; at this point, the hydraulic conductivity approaches
zero. Hence, in the vadose zone, hydraulic conductivity is
expressed as a function of moisture content or pressure head.
Measuring vadose-zone hydraulic conductivity values is
difficult because head gradients, flow rates, and moisture
content or pressure head also must be measured. Factors that
influence these measurements include soil texture, soil struc-
ture, initial water content, depth of water table, water tempera-
ture, entrapped air, biological activity, entrained sediment in
the applied water, and chemistry of the applied water (Wilson,
1982).
Relative permeability also must be determined. The rela-
tive permeability is a normalized coefficient, which when
multiplied by the saturated hydraulic conductivity, yields the
vadose-zone hydraulic conductivity. It is typically presented
as either a function of capillary pressure or saturation. Rela-
tive permeability ranges from one at 100 percent saturation to
zero at residual saturation, the water saturation where the
water phase becomes disconnected and ceases to flow.
A number of empirical equations have been developed
for approximating the vadose-zone permeability of isotropic
porous media. Three commonly used equations for estimating
the vadose-zone hydraulic conductivity are those by Brooks
and Corey (1964), Mualem (1976), and van Genuchten (1980).
Methods to determine the vadose-zone hydraulic conductivity
are listed in Table 5-7 and discussed in Thompson et al.
(1989). Figure 5-4 shows typical relative permeability curves
computed using van Genuchten (1980).
5.2.5 Soil Gas Analysis
Although not strictly flow related, soil gas analysis is an
important remote sensing tool for locating areas contaminated
by VOCs in the vadose zone. This method requires the drilling
of a shallow hole or the injection of a sample tube into the soil.
Volumes of soil gas are evacuated to the surface for collection
and analysis at a remote lab or measured on site by a lab-
quality vapor analyzer. This method also can be used to
analyze cuttings from well drilling operations or in cases
where installed wells yield no water. Soil gas analysis is
dependent upon the pore spacing within the soil and is less
reliable in tightly packed soils such as clay. It also cannot be
used to detect nonvolatile organic compounds and inorganic
compounds (see Table 5-8). Section 9.2.2 provides some
further discussion of soil gas sampling techniques.
Using soil vapor monitoring wells to detect plumes of
ground water contaminated with VOCs has been suggested as
a cost-effective means of tracing ground-water contamination
(e.g., Marrin and Kerfoot, 1988). Indeed, some success in
using this technique has been reported (Marrin and Thomp-
64
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Wetting
Saturation •
Water Content
Figure 5-3. Soil-water characteristic curve displaying hysteresis (modified from Hillei, 1980).
Table 5-6. Summary of Methods for Determining Moisture Characteristic Curves
Method
Porous plate
Vapor equilibration
Osmotic
Application Reference
Standard laboratory method for measurement of soils. Klute (1986)
Can be used to characterize both wetting and drying behavior.
Best suited for matric potentials less than -15 bars. Klute (1986)
Similar to porous plate method. Requires long equilibration times. Klute (1986)
Not recommend.
From Thompson et a/., 7989
Copyright© 1989 Electric Power Research Institute. EPRI EN-6637. Techniques to Develop Data for Hydrogeochemical Models, Reprinted with
permission.
son, 1987; Marrin and Kerfoot 1988). See, also, the soil gas
sampling case studies summarized in Table 9-6.
If the source of the VOCs is below the water table, then
the maximum concentration of the organics in the unsaturated
zone is the top of the capillary fringe. Once the contaminants
have reached the top of the capillary fringe they should diffuse
very rapidly because of the large gas-phase diffusion coeffi-
cients in the unsaturated zone. This rapid mass transfer from
the water in capillary fringe to the soil air just above it should
deplete the capillary fringe of the volatile contaminant. The
concentrations in the unsaturated zone, therefore, are more
controlled by the rate of mass transfer from the ground water
to the top of the capillary fringe, a process controlled by the
very low solute diffusion coefficients. Laboratory studies of
mass transfer across the capillary fringe substantiate these
ideas. With the additional loss of mass by mass transfer across
the atmosphere-soil boundary and by biodegradation that also
may be occurring (e.g., Huh et al, 1987), concentrations in
the unsaturated zone are expected to be very low. The best
opportunity for detecting VOC contaminants under these con-
ditions is to use soil-gas monitoring wells installed just above
the capillary fringe.
There are, of course, exceptions to this scenario. If there
is residual nonaqueous phase liquid (NAPL) in the unsatur-
ated zone or product floating on the water table, then soil gas
monitoring would detect the volatiles. In the absence of any
NAPL, VOCs may be detected by soil-gas monitoring if the
water table fluctuates enough to bring the contaminated water
up into the unsaturated zone and leave it there as part of the
residual phase. The VOCs would then partition from the
65
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Table 5-7.
Method
Constant-head
borehole
infiltration
Guelph
permeameter
Air-entry
permeameter
Instantaneous
profile
Summary of Methods to Measure Vadose-Zona Hydraulic-Conductivity Values in the Field and Laboratory
Application Reference
Field method in open or partially cased borehole. Most commonly used
method. Includes a relatively large volume of porous media in test.
Amoozegar and Warrick (1986)
Field method in open, small-diameter borehole (> 5cm). Relatively fast
method (5 to 60 minutes) requiring small volume of water. K,, K (
and sorptivity are measured simultaneously. Many boreholes and tests
may be required to fully represent heterogeneities of porous media.
Field method. Test performed in cylinder which is driven into porous
media. Small volume of material tested; hence, many tests maybe
needed. Fast, simple method requiring little water (-10 L).
Field or lab method. Field method measures vertical during
drainage. Measurement of moisture content and hydraulic head
needs to be rapid and nondestructive to sample. Commonly used
method, reasonably accurate.
Bouwer (1978);
Stephens and Neuman (1982 a,b,c);
Reynolds and Elrick (1986)
Bouwer (1966)
Bouma et al. (1974):
Klute and Dirksen, (1986)
Crust-imposed
steady flux
Sprinkler-imposed
steady flux
Parameter
identification
Empirical equations
Field method. Measures vertical K( ) during wetting portion of
hysteresis loop. Labor and time intensive.
Field method. Larger sample area than for crust method. Useful
only for relatively high moisture contents.
Results of one field or lab test are used by a numerical approximation
method to develop K(d), KM. and v(0) over a wide range of0 and
Relatively fast method; however, unique solutions are not
usually attained.
Each empirical equation has its own application based upon the
assumptions of the equation. Relatively fast technique.
Green, et al. (1986)
Green, Ahuja, and Chong (1986)
Zachmann et al. (198la,b, 1982);
Kool et al. (1985)
Brooks and Corey (1964);
van Genuchten (1980); Mualem
(1976)
From Thompson et al, 1989
Copyright® 1989 Electric Power Research Institute. EPRI EN-6637. Techniques to Develop Data for Hydrogeochemical Models. Reprinted with
permission.
residual phase into the gas phase where they could be de-
tected.
Given the current understanding of the magnitude of the
processes controlling the rate of migration of organic con-
taminants in the gas phase, it may be more reasonable to
reverse the argument. If there is some NAPL in the unsatur-
ated zone, VOCs can travel significant distances in the gas
phase. Provided that the Henry's constants for these organic
contaminants are sufficiently small, these volatiles can parti-
tion into the infiltrating water and be carried to the subsurface
to form a shallow contaminant plume. So, the ground-water
contaminant plume results from the soil-gas contamination
rather than from the ground water.
5.5 Analysis of Data
There are several programs used to evaluate flow in the
vadose zone, many of which are discussed in van der Heijde et
al. (1988). Because of the nonlinear and hysteretic behavior of
various parameters, modeling vadose-zone flow is more diffi-
cult than modeling saturated flow. There are additional prob-
lems because of the atmospheric boundary conditions
associated with seepage faces, infiltration, and evapotranspi-
ration. Because of the research associated with pesticides,
several programs that analyze the vadose zone are available
through the Center for Exposure Assessment Modeling in
Athens, Georgia. Other vadose-zone programs are available
from the Robert S. Kerr Environmental Research Laboratory
in Ada, Oklahoma, and the International Ground Water Mod-
eling Center, Holcomb Research Institute, Butler University,
Indianapolis, Indiana.
If this describes the interaction between contaminated
soil gas and contaminated ground water, then the greatest use
of shallow soil-gas monitoring surveys is for locating poten-
tial residuals of NAPL in the subsurface. The areas with the
highest gas-phase concentrations are most likely to be those
closest to any residual product. Thus, such a survey could be
an effective guide for determining the optimal locations for
soil-gas extraction wells.
5.4 Remedial Actions
When the vadose zone is shallow, excavation as a reme-
dial action is commonly considered. (An example of excava-
tion with fixation is given in Section 5.5.). For volatile
chemicals near or above the water table, vacuum extinction is
another technique that can remove contaminants from the
residual phase. During vacuum extraction, air is pulled through
66
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Clay
Clay Loam
Loam
Loamy Sand
Silt Loam
X----X
*---•*
Siltv Clay Loam B - - -Q
Sand
Sandy Clay
Sandy Clay Loam Y- - - -Y
Sandv Loam
0.1 -
0.0
0.2
0.3
0.4
0.5 0.6
Water Saturation
Figure 5-4. Relative permeability curves for various soil types.
soils contaminated with VOCs. The resulting vapors move
through the soil and are collected at extraction wells. Simple
techniques that have been developed to control subsurface
hydrocarbon vapors are discussed in O'Connor et al, (1984),
Dunlap (1984), and Marley and Hoag (1984). In general, two
principal types of vapor management systems are available.
The frost type, a positive differential pressure system, induces
vapor flow away from the control points, while the second
type, a negative differential pressure system, induces vapor
flow toward the control points. The vapor management meth-
ods may be either passive or active. Passive methods use
naturally occurring differences in vapor pressures to induce
the required flow regime. Active methods require the artificial
generation of differential vapor pressures to accomplish the
same flow pattern. Practical experience demonstrates that
active generation of negative differential vapor pressures typi-
cally provides the most favorable field results.
The air flow generates advective vapor fluxes that change
the vapor-liquid equilibrium, inducing volatilization of con-
taminants. This method is advantageous because it is imple-
mented in place, and, therefore, causes minimum disruption.
This is especially important at active facilities or sites where
investigations are hindered by physical obstacles. Vacuum
extraction laboratory studies are descriked in Marley and
Hoag (1984), Thornton and Wootan (1982), and Texas Re-
search Institute (1984). Crow et al. (1985, 1987) discusses a
field-scale experiment. Agrelot et al. (1985), Regalbuto et al.
(1988), Connor (1988), and Hutzler et al. (1989) show appli-
cations to hazardous waste sites.
Vacuum extraction can effectively remove chemicals from
the vadose zone. According to Hutzler et al. (1989), most
chemicals successfully extracted have a low molecular weight
and high volatility. Most of the compounds have values of
Henry's Law constants water than 0.01. If the water table is
lowered, vacuum extraction also can be used to remove re-
sidual NAPL from below the original water table elevation.
For example, ground-water pumping and vacuum extraction
are being used together to clean up DNAPL contamination at
the Tyson's Superfund site (Wassersug, 1989). Vacuum ex-
traction also can increase natural biodegradation processes by
introducing additional oxygen into the subsurface. Finally,
vacuum extraction generally is used in conjunction with other
remedial methods.
67
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Table 5-8. Characteristics of Contaminants in Relation to Soil Gas Surveying
Group/Contaminants Applicability of Soil-Gas Survey Techniques
Group A: Halogenated Methanes, Ethanes, and Ethenes
chloroform, vinyl chloride, Detectable in soil gas over a wide range of environmental conditions. Dense non-aqueous phase
carbon tetrachloride, liquid (DNAPL), wi// sink in aquifer if present as pure liquid.
trichlorofluoromethane,
TCA, EDB, TCE
Group B: Halogenated Propanes, Propenes and Benzenes
chlorobenzene, Limited value; detectable by soil-gas techniques oniy where probes can sample near contaminated soil or
trichlorobenzene, ground water. DNAPL.
1,2-dichloropropane
Group C: Halogenated Polycyclic Aromatics
aldrin, DDT Do not partition into the gas phase adequatelv to be detected in soil gas under normal circumstances.
chlordane, heptachlor, DNAPL
PCBs
Group D: C,- CaPetroleum Hydrocarbons
benzene, toluene, Most predictably detected in shallow aquifers or leaking underground storage tanks where probes can be
xylene isomers, methane, driven near the source of contamination. Light nonaqueous phase liquids (LNAPLs), float as thin film on
ethane, cyclohexane, the water table. Can act as a solvent for DNAPLs, keeping them nearer the ground surface.
gasoline, JP-4
Group E: Cg- C,2Petroleum Hydrocarbons
trimelhylbenzene, Limited value; detectable by soil gas techniques only where probes can sample near contaminated
naphthalene, decane, soil or ground water. DNAPL.
diesel and jet A fuels
Group F: Polycylic Aromatic Hydrocarbons
anthracene, benzopyrene, Do not partition adequately into the gas phase to be detected in soil gas under normal circumstances.
fluoranthene, chrysene, DNAPL
motor oils, coal tars
Group G: Low Molecular Weight Oxygenated Compounds
acetone, ethanol, LNAPLs, but dissolve readily in ground water. May be detected in soil gas if they result from a
formaldehyde, leak or spill in relatively dry soil.
methylethylketone
Source: Adapted from Marrin (1987)
5 5 Examvle-Pevver's Steel Site Mer reviewing several remedial options, investigators se-
^ ^^ lected solidification/stabilization. In accordance with regula-
Fixation technology is demonstrated in a case study of the tionS) PCB-contammated oils were removed and disposed of
30-acre Pepper's Steel cleanup site, located near Miami and at an approved facility off site. All the contaminated soils
Medley, Florida, where the Miami Canal borders the site were solidified on site with a proportioned mix of fly ash and
(Figure 5-5). Ground water in the Biscayne aquifer is about 5 cement Solidification changes the physical characteristics of
to 6 ft below land surface. Soils above the aquifer were the waste and decreases the surface area of pollutants avail-
contaminated as a result of prior business operations at the able for ieaching. Through stabilization, the wastes become
site, and polychlormated biphenyls (PCBs) and heavy metals iess water soiubie and less toxic. The PCBs are trapped in the
(lead, arsenic) were found in concentrations significant to cement mixture and the heavy metals (arsenic and lead)
warrant action. become insoluble metal silicates.
The two primary goals of site cleanup were The amount of S0ll excavated for fixation was minimized
. . by using kriging on soil chemistry data. The kriged results
• Collect and dispose of oils containing PCBs that are mdlcated zones of contamination as well as a measure of the
uncovered dunng site excavations. error of estimation. Some details of the cleanup include (U.S.
EPA, 1987):
Treat or dispose of soils that are contaminated with
PCBs and heavy metals. . Approximately 60,000 cubic yards of contaminated
soils were excavated.
68
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} Monolith Perimeter
\ ^
r" gg^3
L ,
1
\
j
— 1 ,'
MW-5B -
** MW-5A Q 50 100 Fee
i
9 MO-3 r—H—^ i
J"^ZZZfli. >
v>^*.~.»_/
OMW-3^ ®MlV-2/A MW-1AQ
pi
a n HH
n 0a °
!
MW-4S
Lftn]fv-4C ^f* j. A/f \A/-dA
Gfy ® * r* * * "**^
Figure 5-5. Location of Pepper's Steel and Alloy site monitoring wells.
• All free oil uncovered during excavation was col-
lected and sent for treatment or disposal.
Soils contaminated with PCBs and heavy metals
were stabilized solidified with a cement mixture.
• Solidified materials were placed back on the Pepper's
Steel site and covered with 12 in. of crushed lime-
stone.
Surface water was controlled by grading the site and
placing drains around the solidified material.
Ground water is monitored annually.
5.6 References
Agrelot, J.C., J.J. Malot, and M.J. Visser. 1985. Vacuum:
Defense System for Ground Water VOC Contamination.
In: Fifth National Symposium and Exposition on Aquifer
Restoration and Ground Water Monitoring, National Water
Well Association, Dublin, OH, pp. 485-494.
Amoozegar, A. and A.W. Warrick. 1986. Hydraulic Conduc-
tivity of Saturated Soils: Field Methods. In: Methods of
Soil Analysis, Part 1, 2nd ed., A. Klute (ed.), Agronomy
Monograph No. 9, American Society of Agronomy, Madi-
son, WI, pp. 735-770.
Bouma, J., F.G. Baker, andP.L.M. Veneman. 1974. Measure-
ment of Water Movement in Soil Pedons above the Water
Table. University of Wisconsin-Extension, Geological
and Natural History Survey, Information Circular No. 27.
Bouwer, H. 1966. Rapid Field Measurement Air Entry Value
and Hydraulic Conductivity of Soil as Significant Param-
eters in Flow System Analysis. Water Resources Re-
search 2(4):729-738.
Bouwer, H. 1978. Ground-Water Hydrology. McGraw-Hill,
New York.
Bouwer, H. 1986. Intake Rate: Cylinder Infiltrometer. In:
Methods of Soil Analysis, Part 1, 2nd ed., A. Klute (ed.),
Agronomy Monograph No. 9, American Society of
Agronomy, Madison, WI, pp. 825-844.
Brooks, R.H. and A.T. Corey. 1964. Hydraulic Properties of
Porous Media. Hydrology Paper No. 3. Colorado State
University, Fort Collins, CO.
Conner, J.R. 1988. Case Study of Soil Venting. Pollution
Engineering 20(7):75-78.
69
-------
Crow, W.L., E.P. Anderson, andE.M. Mmugh. 1985. Subsur-
face Venting of Hydrocarbon Vapors from an Under-
ground Aquifer. In: Proc. NWWA/API Conf. on Petroleum
Hydrocarbons and Organic Chemicals in Ground Wa-
ter-prevention, Detection and Restoration, National
Water Well Association, Dublin, OH, pp. 536-554.
Crow, W.L., E.R. Anderson, and E. Minugh. 1987. Subsur-
face Venting of Hydrocarbons Emanating from Hydro-
carbon Product on Groundwater. Ground Water
Monitoring Review 7(1)51-57.
Davis, S.N. and R.J.M. DeWiest. 1966. Hydrogeology. John
Wiley & Sons, New York.
Dunlap, L.E. 1984. Abatement of Hydrocarbon Vapors in
Buildings. In: Proc. NWWA/API Conf. on Petroleum
Hydrocarbons and Organic Chemicals in Ground Wa-
ter—prevention, Detection and Restoration, National
Water Well Association, Dublin, OH, pp. 504-518.
Dunne, T. and L.B. Leopold. 1978. Water in Environmental
Planning. W.H. Freeman, San Francisco,818 pp.
Finkelstein, P.L., D.A. Mozzarella, T.A. Lockhart, WJ. King,
and J.H. White. 1989. Quality Assurance Handbook for
Air Pollution Measurement Systems, IV: Meteorological
Measurements, revised. EPA/600/4-90-003.
Gardner, W.H. 1986. Water Content. In: Methods of Soil
Analysis, Part 1, 2nd ed., A. Klute (ed.), Agronomy
Monograph No. 9, American Society of Agronomy, Madi-
son, WI, pp. 493-544.
Green, RE., L.R Ahuja, and S.K. Chong. 1986. Hydraulic
Conductivity, Diffusivity, and Sorptiviry of Unsaturated
Soils: Field Methods. In: Methods of Soil Analysis, Part
1, 2nd ed., A. Klute (ed.), Agronomy Monograph No. 9,
American Society of Agronomy, Madison, WI, pp. 771-
798.
Green, W.H. and C.A. Ampt. 1911. Studies on Soil Physics, I:
Flow of Air and Water through Soils. J. Agricultural
Science 4:1-24.
Hillel, D. 1980. Fundamentals of Soil Physics. Academic
Press, New York, 413 pp.
Hult, M.F. and R.R. Grabbe. 1985. Permanent Gases and
Hydrocarbon Vapors in the Unsaturated Zone. In: Pro-
ceedings, U.S. Geological Survey Second Toxic Waste
Technical Meeting, Cape Cod, MA, October 1985.
Hutzler, N.F., B.E. Murphy, and J.S. Gierke. 1989. State of
Technology Review Soil Vapor Extraction Systems. U.S.
EPA Cooperative Agreement CR-8143 19-01-1 (NTIS
PB89-195184), 36 pp.
Klute, A. 1986. Water Retention: Laboratory Methods. In:
Methods of Soil Analysis, Part 1, 2nd ed., A. Klute (ed.),
Agronomy Monograph No. 9, American Society of
Agronomy, Madison, WI, pp. 635-662.
Klute, A. and C. Dirksen. 1986. Hydraulic Conductivity and
Diffusivity: Laboratory Methods. In: Methods of Soil
Analysis, Part 1, 2nd ed., A. Klute (ed.), Agronomy
Monograph No. 9, American Society of Agronomy, Madi-
son, WI, pp. 687-734.
Kool, J.B., J.C. Parker, andM.T. van Genuchten. 1985. Deter-
mining Soil Hydraulic Properties from One-Step Outflow
Experiments by Parameter Estimation: I. Theory and
Numerical Studies. Soil Sci. Soc. Am. J. 49:1348-1354.
Marley, M.C. and G.E. Hoag. 1984. Induced Soil Venting for
Recovery/ Restoration of Gasoline Hydrocarbons in the
Vadose Zone. In: Proc. NWWA/API Conf. on Petroleum
Hydrocarbons and Organic Chemicals in Ground Wa-
ter—prevention, Detection and Restoration, National
Water Well Association, Dublin, OH, pp. 473-503.
Marrin, D.L. 1987. Soil Gas Sampling Strategies: Deep vs.
Shallow Aquifers. In: Proc. 1st Nat. Outdoor Action
Conf. on Aquifer Restoration, Ground Water Monitoring
and Geophysical Methods, National Water Well Associa-
tion, Dublin, OH, pp. 437-454.
Marrin, D.L. and H.B. Kerfoot. 1988. Soil-gas Surveying
Techniques. Environ. Sci. Technol. 22(7):740-745.
Marrin, D.L. and G.M. Thompson. 1987. Gaseous Behavior
of TCE Overlying a Contaminated Aquifer. Groundwater
25:21-27.
Mualem, Y. 1976. A New Model for Predicting the Hydraulic
Conductivity of Unsaturated Porous Media. Water Re-
sources Research 12:593-622.
Musgrave, G.W. and H.N. Holtan. 1964. Infiltration. In: Hand-
book of Applied Hydrology, V.T. Chow (ed.), McGraw-
Hill, New York, pp. 12-1 to 12-30.
National Weather Service. 1972. Observing Handbook No. 2.
Data Acquisition Division, Office of Meteorological Op
emtions, Silver Spring, MD.
O'Connor, M.J., J.G. Agar, and RD. King. 1984. Practical
Experience in the Management of Hydrocarbon Vapors
in the Subsurface. In: Proc. NWWA/API Conf. on Petro-
leum Hydrocarbons and Organic Chemicals in Ground
Water-Prevention, Detection and Restoration, National
Water Well Association, Dublin, OH, pp. 519-533.
Peterson, A.E. and G.D. Bubenzer. 1986. Intake Rate Sprin-
kler Infiltrometer. In: Methods of Soil Analysis, Part 1,
2nd ed., A. Klute (ed.), Agronomy Monograph No. 9,
American Society of Agronomy, Madison, WI, pp. 845-
870.
Philip, J.R. 1957. The Theory of Infiltration, I: The Infiltration
Equation and its Solution. J. Soil Science 83:345-357.
Radian Corporation. 1988. FGD Chemistry and Analytical
Methods Handbook, 2 Chemical and Physical Test Meth-
70
-------
ods, Revision 1. EPRI CS-3612. Electric Power Research
Institute, Palo Alto, CA. [Originally published in 1984].
Regalbuto, D.P., J.A. Barrera, and J.B. Lisiecki. 1988. In-situ
Removal of VOCs by Means of Enhanced Volatilization.
In: Proc. NWWA/API Conf. on Petroleum Hydrocarbons
and Organic Chemicals in Ground Water-Prevention,
Detection and Restoration, National Water Well Associa-
tion, Dublin, OH, pp. 571-590.
Reynolds, W.D. and D.E. Elrick. 1986. A Method for Simul-
taneous In Situ Measurement in the Vadose Zone of Field
Saturated Hydraulic Conductivity, Sorptivity and the Con-
ductivity-pressure Head Relationship. Ground Water
Monitoring Review 6(4):84-95.
Schmugge, T.J., T.J. Jackson, and H.L. McKim. 1980. Survey
of Methods for Soil Moisture Determination. Water Re-
sources Research 16(6):961-979.
Sharma, M.L. 1985. Estimating Evapotranspiration. In: Ad-
vances in Irrigation, 3. Academic Press, New York.
Stephens, D.B. and S.P. Neuman. 1982a. Vadose Zone Per-
meability Tests: Summary. J. Hydraulics Division ASCE
198(HY5):623-639.
Stephens, D.B. and S.P. Neuman. 1982b. Vadose Zone Per-
meability: Steady State Results. J. Hydraulics Division
ASCE 198(HY5):640-659.
Stephens, D.B. and S.P. Neuman. 1982c. Vadose Zone Per-
meability Unsteady Flow. J. Hydraulics Division ASCE
198(HY5): 660-677.
Stephens, D.B. and S. Heermann. 1988. Dependence of Anisot-
ropy on Saturation in a Stratified Sand. Water Resources
Research 24(5):770-778.
Stephens, J.C. and E.H. Stewart. 1964. A Comparison of
Procedures for Computing Evaporation and Evapotrans-
piration. Agricultural Research Service, Ft. Lauderdale,
FL.
Texas Research Institute. 1984. Forced Venting to Remove
Gasoline Vapors from a Large-Scale Model Aquifer.
American Petroleum Institute, Washington, DC, 60 pp.
Thompson, C.M., et al. 1989. Techniques to Develop Data for
Hydrogeochemical Models. EPRI EN-6637. Electric
Power Research Institute, Palo Alto, CA.
Thomthwaite, C.W. and J.R. Mather. 1957. Instructions and
Tables for Computing Potential Evapotranspiration and
Water Balance. Drexel Institute of Technology, Labora-
tory of Climatology, X(3).
Thornton, J.S. and W.L. Wootan. 1982. Venting for the Re-
moval of Hydrocarbon Vapors from Gasoline Contami-
nated Soil. J. Environmental Science and Health
A17(l):31-44.
U.S. Environmental Protection Agency. 1987. Protecting the
Biscayne Aquifers: Actions to be Taken at the Pepper's
Steel and Alloy Site. Prepared by CH2M Hill.
U.S. Geological Survey. 1977. National Handbook of Recom-
mended Methods for Water Data Acquisition (Chapter
8—Evaporation and Transpiration, updated June 1982).
USGS Office of Water Data Coordination, Reston, VA.
van Bavel, C.H.M. 1963. Soil Moisture Measurement with the
Neutron Method. USDA-ARS, ARS-41-70, U.S. Gov-
ernment Printing Office, Washington, DC.
van Genuchten, M.T. 1980. A Closed Form Equation for
Predicting the Hydraulic Conductivity of Unsaturated
Soils. Soil Sci. Soc. Am. J. 44:892-898.
van der Heijde, P. K. M., A.I. El-Kadi, and S.A. Williams.
1988. Groundwater Modeling: An Overview and Status
Report. EPA/600/2-89/028 (NTIS PB89-224497). Also
available from International Ground Water Modeling
Center, Butler University, Indianapolis, IN.
Veihmeyer, F.J. 1964. Evapotranspiration. In: Handbook of
Applied Hydrology, V.T. Chow (ed.), McGraw-Hill, New
York, pp. 11-1 to 11-38.
Wassersug, S.R. 1989. Policy Aspects of Current Practices
and Applications. In: Remediating Groundwater and Soil
Contamination, Report on a Colloquium, Water Science
and Technology Board, National Academy Press, Wash-
ington, DC.
Wilson, L.G. 1980. Monitoring in the Vadose Zone: A Re-
view of Technical Elements and Methods. EPA/600/7-
80-134 (NTIS PB81-125817), 168 pp.
Wilson, L.G. 1981. Monitoring in the Vadose Zone, Part I:
Storage Changes. Ground Water Monitoring Review
1 (3): 32-41.
Wilson, L.G. 1982. Monitoring in the Vadose Zone, Part II:
Ground Water Monitoring Review 2(4):31-42.
Wilson, L.G. 1983. Monitoring in the Vadose Zone, Part III:
Ground Water Monitoring Review 3(4): 155-166.
Zachmann, D.W., P.C. DuChateau, and A. Klute. 1981a. The
Calibration of the Richards Flow Equation for a Draining
Column by Parameter Identification. Soil Sci. Soc. Am. J.
45:1012-1015.
Zachmann, D.W., P.C. DuChateau, and A. Klute. 1981b. The
Estimation of Soil Hydraulic properties from Inflow Data.
In: Proceedings, Symposium on Rainfall-Runoff Model-
ing, V.V. Singh (ed.), Water Resources Publications,
Littleton, CO, pp. 173-180.
Zachmarm, D.W., P.C. DuChateau, and A. Klute. 1982. Si-
multaneous Approximation of Water Capacity and Soil
Conductivity by Parameter Identification. Soil Science
134:157-163.
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Chapter 6
Characterization of Water Movement in Saturated Fractured Media
James W. Mercer and Charles P. Spalding
Characterizing heterogeneity and anisotropy in the sub-
surface is important, especially in fractured or karst media.
Fracturing or caverns provide preferential flow paths for
ground water. Many of the characterization tools and tech-
niques discussed for porous media also may be used for
fractured media, if care is used to interpret the data. Tech-
niques that are particularly helpful in understanding fractured/
cavernous media include coring, aquifer tests, tracer tests,
geophysical tools, geochemical techniques, and fracture trace
analysis. Most of these techniques are discussed in this chap-
ter.
As in the preceding chapters, the discussion begins with a
review of concepts. This review is followed by sections on
field techniques, analysis of data, and a case study. This
chapter draws freely upon material contained in a recent EPA
Superfund ground water-issue paper on contaminant transport
in fractured media (Schmelling and Ross, 1989).
6.1 Review of Concepts
Most fractured bedrock systems consist of rock bounded
by discrete discontinuities composed of fractures, joints, and
shear zones, usually occurring in sets with similar geometries
(Witherspoon et al, 1987). Figure 6-1 illustrates this type of
Shear Zone
Joints
Solid Rock
Fracture Zone
Figure 6-1.
Conceptualization of discontinuities in a fractured
medium (from Gale, 1982).
system, referred to as a dual-porosity system. In addition to
the discontinuities shown in the figure, bedding planes also
can behave as discontinuities. Fractures may be open, min-
eral-filled, deformed, or any combination thereof (Nelson,
1985).
Open fractures may provide conduits for ground-water
and contaminant movement through a rock mass that is other-
wise relatively impermeable. Fractures may be filled either
partially or completely by secondary cementing materials.
such as quartz or carbonate minerals, which reduce or elimi-
nate fracture porosity and permeability. The permeability of
deformed fractures also may be reduced by gouge, a finely
abraded material produced by the cataclasis of grains in
contact across a fault plane during displacement of the rock
masses. Slickensides, striated surfaces formed by frictional
sliding along a fault plane, also are a deformed-fracture fea-
ture. Slickensides reduce permeability perpendicular to the
fracture plane, but the mismatch of fracture surfaces may
increase permeability along the fracture plane. Very little
displacement is necessary to produce gouge or slickensides.
Another factor that may reduce permeability is the deposition
of a thin layer of low- permeability material called a fracture
skin. This skin prevents the free exchange of fluids between
the rock matrix and the fracture (Moench, 1984).
The concept of fracturing presented so far is one element
of a more complicated hierarchy of multiple-porosity systems.
In soluble bedrock like limestone, dolostone, or evaporates,
conduit flow can develop as original fracture systems are
enlarged by solution. The important feature of conduit flow,
when it is able to develop, is the integration of the drainage
network (Quinlan and Ewers, 1985). In many ways, the net-
work is analogous to a river system with smaller tributaries
supplying water to a succession of larger and larger conduits.
As a result of the integration, both the conduit system and the
individual conduits can become large. For example, the karst
system at Mammoth Cave, Kentucky, has over 330 miles of
connected passages.
Major factors affecting ground-water flow through frac-
tured rock include (1) fracture density, (2) orientation, (3)
effective aperture width, and (4) the nature of the rock matrix.
Fracture density (number of fractures per unit volume of rock)
and orientation are important determinants of the degree of
interconnection of fracture sets, which is a critical feature
contributing to the hydraulic conductivity of a fractured rock
73
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system (Witherspoon et al., 1987). Only interconnected frac-
tures provide pathways for ground-water flow and contami-
nant transport, although the flow network may be a subset of
the fracture network. Fractures oriented parallel to the hydrau-
lic gradient are more likely to provide effective pathways than
fractures oriented perpendicular to the hydraulic gradient.
Flow in fractured rock systems can be similar to flow in
porous media when (1) the fracture apertures are constant, (2)
the fracture orientations are randomly distributed, and (3) the
fracture spacing is small relative to the scale of the system
(Longetal, 1982).
The cross-sectional area of a fracture will have an impor-
tant effect on flow through the fracture. Under certain condi-
tions, fracture-flux is generally proportional to the cube of the
fracture aperture (distance between rock blocks) when aper-
tures exceed 10 microns (Witherspoon et al., 1987). Fracture
apertures and, therefore, flow through fractures are highly
stress-dependent and generally decrease with depth (Gale,
1982).
The nature of the rock matrix affects the movement of
water and contaminants through fractured rock systems. Meta-
morphic and igneous rocks generally have very low primary
porosity and permeability. Fractures may account for most of
the permeability in such systems, and the movement of water
and contaminants into and out of the rock matrix may be
minimal. On the other hand, sedimentary rocks generally have
higher primary porosity and varying permeability. Examples
include coarse-grained materials, such as sandstone, which
have relatively high primary porosity and significant matrix
permeability, and fine-grained materials, such as shale, which
have high primary porosity and low permeability.
Fractures may enhance the permeability of all types of
materials. High porosity allows significant storage of water
and contaminants in the rock matrix. Authigenic clays formed
during the weathering on certain rock-forming minerals may
significantly reduce the porosity and permeability of the frac-
tures and rock matrix. Rates of contaminant migration into
and out of the rock matrix will depend on the permeability of
the matrix, the presence of low-permeability fracture skins,
and the matrix diffusion coefficient of the contaminant (Fig-
ure 6-2).
A complete description of a contaminated fractured rock
system would include data on (1) the dimensions of the
system; (2) individual fracture length, aperture width, loca-
tion, and orientation; (3) the hydraulic head throughout the
system; (4) the porosity and permeability of the rock matrix;
(5) the sources of water and contaminants; (6) the nature and
concentrations of the contaminants throughout the system;
and (7) the chemical interactions between the contaminants
and rock matrix. Presently, collection of such detailed infor-
mation is neither technically possible nor economically fea-
sible at the scale of most contaminated sites.
6.2 Field Techniques
Hydrogeologic characterization methods usually are most
successful when used in conjunction with one another. These
methods may include coring, aquifer tests, tracer tests, surface
Fracture
Flow
Figure 6-2. Flow through fractures and diffusion of contami-
nants from fractures into the rock matrix of a dual-
porosity medium (from Anderson, 1984).
and borehole geophysical techniques, and use of borehole
flowmeters, or other tools. Important information may be
gathered before, during, and after drilling operations.
6.2.1 Fracture Trace Analysis
Ground-water flow in bedrock is generally concentrated
in the upper weathered zone of the rock and in fractures at
depth. A well penetrating a zone of subsurface fractures,
therefore will yield more water than a well drilled in an area
with relatively few fractures. Such zones are also pathways
for contaminant migration. Selecting drill sites by examining
aerial photographs stereoscopically for surficial expressions
of linear zones of subsurface fractures will increase the prob-
ability of high yields and locating contaminants. This type of
study is known as fracture trace analysis (Ray, 1960; Fetter,
1980). Figure 6-3 shows the relationship between fracture
traces and zones of fractures. In general, higher yields can be
expected in topographic low areas because (1) swales and
valleys tend to be cut into less-resistant, more highly fractured
and more-permeable rock; and (2) ground-water flow usually
converges in stream valleys.
Although fracture traces, fault planes, and other linea-
ments are often identifiable on aerial photographs, they must
be field-verified to distinguish anthropogenic features such as
fences and buried pipelines from geologic features. The orien-
tation of all fractures (e.g., outcrops and excavations) identi-
fied from aerial photographs and field observations should be
measured and plotted on maps as well as on rose diagrams
(where the frequency of fracture orientation is plotted) to
identify major fracture trends. Such trends are usually related
to the geologic (tectonic) history of a site. A basic understand-
ing of a site's tectonic history and subsequent fracture orienta-
tion allows a better understanding of potential contaminant
pathways.
74
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700
200
300
Textural and
Compositional
Variation
I Zone of
. Fracture —
/ Concentration
Figure 6-3. Relationship between fracture traces end zones of fracture concentration (after Lattman and Parizek, 1964).
6.2.2 Coring
Core material obtained during drilling operations can
yield information on the density, location, and orientation of
fractures and provide samples for physical and chemical
testing. Core samples also may provide information concer-
ning fracture roughness and mineral precipitation on fracture
surfaces. Information collected during coring operations with
open hole completions should include (1) the location of
major water-bearing fractures, (2) changes in hydraulic head
with depth, and (3) changes in the ground-water geochemis-
try, Water loss to a fracture zone, drilling rates, and the
presence of contaminants also are useful active drilling data
(this information is discussed in detail in Chapter 4). In certain
instances, cores may be taken diagonally to intercept near
vertical fractures and determine fracture azimuth. While a
major drawback of coring can be the relatively high cost, the
information obtained often makes this characterization tech-
nique cost effective.
6.2.3 Aquifer Tests
Aquifer tests, including constant rate pumping tests and
slug tests, can provide hydraulic conductivity and information
on anisotropy for fractured formations. These tests also allow
the estimation of average fracture apertures of a medium. The
same tests commonly used for unconsolidated porous media
can be used for fractured media, but the test results will
generally be more difficult to interpret. Barker and Black
(1983) note that transmissivity values will always be overesti-
mated by applying standard type curve analysis to fissured
aquifers.
Other more complex tests, such as cross-hole packer
tests, are particularly applicable to fractured media. For ex-
ample, Hsieh and Neuman (1985) and Hsieh et al. (1985)
describe a method of determining the three-dimensional hy-
draulic conductivity tensor. The method consists of injecting
fluid into, or withdrawing fluid out of, selected intervals
isolated by inflatable packers and monitoring the transient
response in isolated intervals of neighboring wells.
This method is applicable to situations where the princi-
pal directions of the hydraulic conductivity tensor are not
necessarily vertical and horizontal. A minimum of six cross-
hole tests is required to determine the six independent comp
nents of the hydraulic conductivity tensor. In practice, scatter
in the data is likely to be such that more than six cross-hole
tests will be required. Hsieh and his coworkers concluded that
failure to fit data to an ellipsoidal representation indicated that
the rock under study could not be represented by an equiva-
lent, continuous, uniform, anisotropic medium the scale of the
test. Depending on the application to be made, the test may be
repeated on a larger scale or the data may be interpreted in
terms of discrete fractures of the system.
While aquifer tests can provide information on aquifer
anisotropy, heterogeneity, and boundary conditions, they do
not provide information on the range of fracture apertures or
75
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surface roughness. One of the major drawbacks associated
with long-term aquifer testing is the necessity to store and
treat the large volume of water discharged during the test.
Results of aquifer tests in fractured media often demon-
strate S-shaped response curves. Early in the pumping test, the
fractures control the yield to the well; therefore, the fracture
properties control the aquifer response. Once the fractures
drain, there is a transition period followed by a time period
during which the porous block properties control the aquifer
response (see Streltsova, 1988).
6.2.4 Tracer Tests
Tracer tests can provide information on effective poros-
ity, dispersion, and matrix diffusion generally unobtainable
from other hydrogeologic methods. Tracer tests either can be
conducted under natural-gradient or forced-gradient condi-
tions. The primary disadvantages of tracer tests are the time,
expense, number of necessary sampling points, and difficul-
ties associated with data interpretation. However, the impor-
tant information provided by tracer tests is difficult to obtain
by any other means. Davis et al. (1985) provide an introduc-
tion to the use of tracers in ground-water investigations (see
also discussion of this report by Quinlan, 1986, and reply by
Davis, 1986). Tracers, most commonly fluorescent dyes, also
are used to help map karst areas (LaMoreaux et al., 1989;
Mull et al., 1988; Quinlan, 1986, 1989).
Graphical geochemical techniques commonly used in
porous media may provide valuable information at fractured
rock sites. Hem (1985) and Lloyd and Heathcote (1985)
provide overviews of methods used to identify the sources and
extent of ground-water mixing. Environmental isotopes, such
as tritium, also are used to interpret pathways and travel times
(LaMoreaux et al., 1989).
6.2.5 Geophysical Tools
Both surface and borehole geophysical methods can be
used to characterize fractured rock systems. Application of
surface geophysical methods such as direct current electrical
resistivity, electromagnetic induction methods, ground-pen-
etrating radar, magnetometer surveys, and seismic and remote
sensing techniques should be evaluated before a drilling pro-
gram is initiated. These techniques may provide insight for
locating potential monitoring wells by identifying the location
of contaminant plumes or the orientation of major fracture
systems. However, the correlation of major surface geophysi-
cal features with contaminant transport processes in fractured
media has yet to be thoroughly characterized.
Borehole walls are usually less susceptible than cores to
fractures induced during drilling operations. Borehole geo-
physical techniques can usually provide a more reliable esti-
mate of fracture density than can cores. However, as indicated
by Nelson (1985) in a review of down-hole techniques, re-
sponses used to detect fractures on well logs are nonunique
and require detailed knowledge of the tool and the various
rock property effects that could cause fracture-like responses.
Borehole geophysical methods include acoustic, electrical
resistivity, caliper, gamma, and other high-energy logging
techniques. The acoustic televiewer presents a continuous
image of the acoustic response of the borehole face and can
detect fracture apertures as small as one millimeter. This
oriented tool also allows the determination of fracture orienta-
tions. Caliper logs are best suited for determining relative
fracture intensity in continuous, competent rock. Advances in
electronic miniaturization have led to the development of
down-hole cameras, capable of providing in situ viewing of
fractures in the subsurface (Morahan and Dorrier, 1984).
6.2.6 Borehole Flowmeters
Flowmeters have been used for many years in industry.
Only recently, however, has instrumentation been developed
that can accurately measure very low flow rates. Borehole
flowmeters measure the incremental discharge along screened
or open-hole portions of wells during small-scale pumping
tests. The three major types of flowmeters currently being
developed are impeller, heat-pulse, and electromagnetic. Heat-
pulse and electromagnetic flowmeters have no moving parts
that may deteriorate over time; they also are more sensitive
than impeller flowmeters (Young and Waldrop, 1989). This
greater sensitivity may allow the detection of the vertical
movement of water within the borehole under nonpumping
conditions. Under pumping conditions, fracture zones con-
tributing ground water to a borehole may be identified.
6.3 Analysis of Data
Flow in fractured media has been modeled using one of
three possible conceptualizations: (1) an equivalent porous
continuum, (2) a discrete fracture network, and (3) a dual-
porosity medium (National Research Council, 1990). The first
of these approaches assumes that the medium is fractured to
the extent that it behaves hydraulically as a porous medium.
The actual existence of fractures is reflected in the choice of
values for the material coefficients (e.g., hydraulic conductiv-
ity, storativity, or relative permeability). Often these param-
eters take on values significantly different from those used for
modeling a porous medium (Shapiro, 1987). Examples of this
approach as cited by Shapiro (1987) are presented in Elkins
(1953), Elkins and Skov (1960), and Grisak and Cherry
(1975).
With the discrete fracture approach, most or all of the
ground water moves through a network of fractures. This
approach assumes that the geometric character of each frac-
ture (e.g., position in space, length, width, and aperture) as
well as the pattern of connection among fractures are known
exactly. In the simplest theoretical treatment, the blocks are
considered to be impermeable. Figure 6-4a is an idealization
of a two-dimensional network of fractures consisting of two
different sets. Note how each fracture, represented on the
figure by a line segment, has a definite position in space,
length, and aperture. The hydraulic characteristics of the
fracture system develop as a consequence of the intersection
of the individual fractures. In three dimensions, the network
can be described in terms of intersecting planes that could be
rectangular (Figure 6-4b) or circular (Figure 6-4c). Examples
of the discrete fracture treatment of flow in networks are
included in Long et al. (1982), Long (1985), Robinson (1984),
Schwartz et al. (1983), and Smith and Schwartz (1984).
76
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The dual-porosity conceptualization of a fractured me-
dium considers the fluid in the fractures and the fluid in the
blocks as separate continua. Unlike in the discrete approaches,
no account is taken of the specific arrangement of fractures
with respect to each other-there is simply a mixing of fluids
in interacting continua (Shapiro, 1987). In the most general
formulation of the dual-porosity model, the possibility exists
for flow through both the blocks and the fractures, with a
transfer function describing the exchange between the two
continua. Thus, a loss in fluid from the fracture represents a
gain in fluid in the blocks (Shapiro, 1987).
Although modeling tools exist to deal with fractured
media, at present, results should be interpreted with caution.
Systems are often complex and extraordinarily difficult to
characterize, especially with the level of effort considered
normal for most site investigations. The state of the art in field
testing provides a relatively rudimentary estimate of values
for some parameters like hydraulic conductivity, while other
parameters, like storativity, must be established through fit-
ting simple theoretical models (usually of the porous medium
type).
6.4 Remedial Actions
In principle, the remedial actions discussed for porous
media apply to fractured media. However, the remediation for
fractured media is usually more difficult to design and imple-
ment. For example, there are two major problems associated
with pump-and-treat technologies: (1) hydraulic conductivity
reduction with stress; and (2) matrix diffusion.
Fractures are difficult to work with because apertures
and, hence, hydraulic conductivity, depend on the stress within
the medium. A fracture can be opened or closed simply by
reducing or increasing the forces applied to it, For example,
pumping a well in a fractured medium reduces the pore
pressure, effectively decreasing the fracture aperture. Gale
(1982) describes a number of empirical-theoretical approaches
designed to model the stress coupling to hydraulic conductiv-
ity.
For heterogeneous conditions such as fractured media,
advected water will sweep through the higher permeable
zones (fractures), removing contamination from those zones.
Movement of contaminants out of the less-permeable zones is
a slower process than advective transport in the higher perme-
ability zones. The contaminants either are slowly exchanged
by diffusion with the flow water present in the larger pores or
move at relatively slower velocities in the smaller pores. A
rule of thumb is that the longer the site has been contaminated
and the more lenticular (layered) the geologic material, the
longer will be the tailing effect. The water and contaminants
residing in the more permeable zones are those first mobilized
during pumping. Thus, pump-and-treat technologies work in
heterogeneous media, but cleanup times will be longer and
more difficult to estimate than for similar systems in more
homogeneous media.
6.5 Example-Marion County, Florida
This example involves site characterization in Marion
County, Florida, at a site located approximately 10 mi west of
^m&
mm
y$$y;!$
Figure 6-4, Three different conceptualizations of fracture networks: (a) a two-dimensional system of line segments (from Shimo
and Long, 1987); (b) a three-dimensional system of rectangular fractures (from Smith et al., 1985); and (c) a three.
dimensional syetem of "penny-shaped" creeks (from Long, 1986).
77
-------
T-15-S
R-20-E
SEC-17
NE1/4-SE1/4-SE1/4
/ Strong
Strength of
Expression
III Weak
0 660 1320 Feet
I—I-H
0 200 400 Meters
Figure 6-5. Fracture-trace expreasions based on photo interpretation.
Ocala. The work, performed for the Southwest Florida Water
Management District (Ward et al., 1989), concerned water
resource assessment of the Floridan aquifer however, many
of the steps and techniques used to characterize the site are
similar to those that would be used at a hazardous waste
facility overlying fractured media. Some of the work is de-
scribed in Giffm and Ward (1989).
The first step of the assessment was to perform a fracture-
trace analysis using aerial photographs. Photolinears were
classified as I, II, or III depending on the strength and continu-
ity of their linear patterns on the photo. Class I photolinears
have the strongest, most continuous expression; Class III have
the weakest. Figure 6-5 shows the fmcture-trace map and the
location of Regional Observation Monitoring Program (ROMP)
Well 120.
After field checking the mapped fractures, the next step
was to confirm them using surface geophysics. The tri-poten-
tial method was used (Ogden and Eddy, 1984; Habberjam,
1969), and the results of this geophysical survey were used to
pinpoint two lineaments within a few hundred feet of a
site where aquifer testing would be performed.
To help locate monitoring wells for the aquifer testing,
numerical modeling was performed using a fracture flow
code. Data typical for that part of Florida were used to
estimate the response to pumping. Based on the field work
and the modeling, the wells were located as shown in Figure
6-6. The locations of some wells were modified due to access
difficulties; three wells were located to penetrate fracture or
solution channel zones; and one well was sited within the
limestone matrix.
After drilling the wells, the investigators performed bore-
hole geophysics tests including caliper, gamma-gamma, and
neutron. In general, cavernous zones are located using the
caliper log, whereas shalely zones that are less likely to form
cavernous zones are located using the gamma-gamma log.
The neutron log is used to indicate porous zones, which
should correspond to caverns. Unfortunately, the geophysical
logs were not useful in differentiating between areas of solu-
tion features (OW1, OW2, and OW3) and rock matrix (OW4).
The final step in the characterization of this site was to
perform hydraulic testing. Both slug tests and an aquifer test
performed at the site demonstrated an underdamped response
(see Figure 6-7). In this type of response, the water level in the
well oscillates due to inertial effects, which are common in
highly permeable aquifers. Vart der Kamp (1976) presents a
method for analyzing underdamped responses to slug tests.
Pumping tests were analyzed using classical Theis analysis
and anew approach based on early-time deviations (Ward and
Giffm, 1989, and Shapiro, 1989).
As a result of site data analysis, dual-porosity
conceptualization, thought to be appropriate for this site, did
not need to be observed in the field testing. The site was used
to develop a regional dual-porosity and discrete fracture model,
which was then calibrated with transient response at wells and
major spring discharges. A conceptual composite of the site
and model response (Figure 6-8) demonstrates the dramatic
difference in the site-scale storage as compared to the re-
gional-scale matrix response. This difference is evidenced by
a four order of magnitude shift in time forming the dual-
porosity envelope.
78
-------
Romp
i
25W
i
0
25E
i
50
OW1 (Proposed)
ii i i
75 100 125 150E
r 25N
- 0
I 25S
Intersection Location
75M
0 *"
17M
South
Edge of Road
Figure 6-6. Location of four observation wells in the vicinity of ROMP 120.
10'
1
Q
to-'
10
-2
10-
OW 1 (Deep) Drawdown
Theis Analysis
* W(u) = 1
1/u = 3.5
S =.11ft
t = .42min
Time (mm)
Periodic Measurement
Q = 180gpm
r = 136ft
* Coordinates of type curve overlay and graph
Figure 6-7. Pump test Interpretation using the deep transducer at monitor well OW1 using Theis method.
79
-------
Local Fracture
Limiting Curve
[^ Oscillatory I
| Response \
Regional Matrix
Limiting Curve
0.01
1.0
10
Time (min)
100
1000
Figure 6-6. Conceptual composite of aquifer test and dual-porosity model response.
80
-------
6.6 References
Anderson, M.P. 1984. Movement of Contaminants in Ground
Water Ground Water Transport-Advection and Disper-
sion. In: Ground-Water Contamination, National Acad-
emy Press, Washington, DC, pp. 37-45.
Barker, J.A. and J.H. Black. 1983. Slug Tests in Fissured
Aquifers. Water Resources Research 19:1558-1564.
Davis, S.N. 1986. Reply to the Discussion by James F. Quinlan
of Ground-Water Tracers. Ground Water 24(3):398-399.
Davis, S.N., D.J. Campbell, H.W. Bentley, and T.J. Flynn.
1985. Introduction to Ground-Water Tracers. EPA 600/2-
85/022 (NTIS PB86-100591). Also published under the
title Ground-Water Tracers in NWWA/EPA Senes, Na-
tional Water Well Association, Dublin, OH.
Elkins, L.F. 1953. Reservoir Performance and Well Spacing,
Spraberry Trend Area Field of West Texas. Trans. Ameri-
can Institute of Mining Engineers 198:177-196.
Elkins, L.F. and A.M. Skov. 1960. Determination of Fracture
Orientation from Pressure Interference. Trans. American
Institute of Mining Engineers 219:301-304.
Fetter, Jr., C.W. 1980. Applied Hydrogeology. Charles E.
Merrill, Columbus, OH, pp. 406-411.
Gale, J.E. 1982. Assessing the Permeability Characteristics of
Fractured Rock. GSA Special Paper 189. Geological
Society of America, Boulder, CO, pp. 163-181.
Giffin, D.A. and D.S. Ward. 1989. Analysis of Early-Time
Oscillatory Aquifer Response. In: Proc. Conf. on New
Field Techniques for Quantifying the Physical and Chemi-
cal Properties of Heterogeneous Aquifers (Dallas, TX),
National Water Well Association, Dublin, OH, pp. 187-
211.
Grisak, G.E. and J.A. Cherry. 1975. Hydrologic Characteris-
tics and Responses of Fractured Till and Clay Confining a
Shallow Aquifer. Canadian Geotechnical Journal 12:23-
43.
Habberjam, G.M. 1969. The Location of Spherical Cavities
Using a Tri-Potential Technique. Geophysics 34(5):780-
784.
Hem, J.D. 1985. Study and Interpretation of the Chemical
Characteristics of Natural Water, 3rd ed. U.S. Geological
Survey Water-Supply Paper 2254,263 pp.
Hsieh, P.A. and S.P. Neuman. 1985. Field Determination of
the Three-Dimensional Hydraulic Conductivity Tensor
of Anisotropic Media, 1. Theory. Water Resources Re-
search 21:1655-1665.
Hsieh, P.A., S.P. Neuman, G.K. Stiles, and E.S. Simpson.
1985. Field Determination of the Three-Dimensional Hy-
draulic Conductivity Tensor of Anisotropic Media, 2.
Methodology and Application to Fractured Rocks. Water
Resources Research 21:1667-1676.
LaMoreaux, P.E., E. Prohic, J. Zoetl, J.M. Tanner, and B.N.
Roche. 1989. Hydrology of Limestone Terranes Anno-
tated Bibliography of Carbonate Rocks, Volume 4. Inter-
national Association of Hydrogeologists, International
Contributions to Hydrogeology, Vol. 10, Verlag Heinz
Heise GmbH, Hannover, Germany, 267 pp.
Lattman, L. H., and R.R. Parizek. 1964. Relationship between
Fracture Traces and the Occurrence of Ground Water in
Carbonate Rocks. J. Hydrology, 2:73-91.
Lloyd, J.W. and J.A. Heathcote. 1985. Natural Inorganic
Hydrochemistry in Relation to Ground Water. Clarendon
Press, Oxford, 296 pp.
Long, J. C. S., J.S. Remer, C.R. Wilson, and P.A. Witherspoon.
1982. Porous Media Equivalents for Networks of Discon-
tinuous Fractures. Water Resources Research 18:645-
658.
Long, J.C.S. 1985. Verification and Characterization of Frac-
tured Rock at AECL Underground Research Laboratory.
BMI/OCRD- 17. Office of Crystalline Repository Devel-
opment, Battelle Memorial Institute, 239 pp.
Long, J. C. S., P. Gilmour, and P.A. Witherspoon. 1985. A
Model for Steady Fluid Flow in Random Three-Dimen-
sional Networks of Disc-Shaped Fractures. Water Re-
sources Research 21:1105-1115.
Moench, A.F. 1984. Double-Porosity Models for a Fissured
Ground Water Reservoir with Fracture Skin. Water Re-
sources Research 20831-846.
Morahan, T. and R.C. Dorrier. 1984. The Application of
Television Borehole Logging to Ground-Water Monitor-
ing Programs. Ground-Water Monitoring Review
4(4): 172-175.
Mull, D.S., T.D. Lieberman, J.L. Smoot, and L.H. Woosely,
Jr. 1988. Application of Dye-Tracing Techniques for
Determining Solute-Transport Characteristics of Ground
Water in Karst Terranes. EPA 904/6-88-001, Region 4,
Atlanta, GA.
National Research Council. 1990. Ground-Water Models:
Scientific and Regulatory Applications. National Acad-
emy Press, Washington, DC, 303 pp.
Nelson, R.A. 1985. Geologic Analysis of Naturally Fractured
Reservoirs. Contributions in Petroleum Geology and En-
gineering, Vol. 1. Gulf Publishing Company, Houston,
TX, 320 pp.
Ogden, A. and P.S. Eddy, Jr. 1984. The Use of Tn-Potential
Resistivity to Locate Fractures and Caves for High Yield
Water Wells. In: NWWA/EPA Conf. on Surface and
Borehole Geophysical Methods in Ground Water Investi-
81
-------
gations (San Antonio, TX), National Water Well Asso-
ciation, Dublin, OH, pp. 130-149.
Quinlan, J.F. 1986. Discussion of "Ground Water Tracers" by
Davis et al. (1985) with Emphasis on Dye Tracing, Espe-
cially in Karst Terranes. Ground Water 24(2):253-259
and 24(3):396-397 (References).
Quinlan, J.F. 1989. Ground-Water Monitoring in Karst Ter-
ranes: Recommended Protocols and Implicit Assump-
tions. EPA 600/X-89/050, EMSL, Las Vegas, NV.
Quinlan, J.F. and R.O. Ewers. 1985. Ground Water Flow in
Limestone Terrains: Strategy Rationale and Procedure
for Reliable, Efficient Monitoring of Ground Water Qual-
ity in Karst Areas. In: Proc. Fifth National Symposium
and Exposition on Aquifer Restoration and Ground Wa-
ter Monitoring, National Water Well Association, Dublin,
OH, pp. 197-234.
Ray, R.G. 1960. Aerial Photographs in Geologic Interpreta-
tion and Mapping. Geological Survey Professional Paper
373,230pp.
Robinson, P.C. 1984. Connectivity Flow and Transport in
Network Models of Fractured Media. DP 1072. Theoreti-
cal Physics Division, AERE, Harwell, U.K.
Schmelling, S.G. and R.R. Ross. 1989. Contaminant Trans-
port in Fractured Media Models for Decision Makers.
EPA Superfund Ground Water Issue Paper. EPA/540/4-
89/004.
Schwartz, F.W., L. Smith, and A.S. Crowe. 1983. A Stochas-
tic Analysis of Macroscopic Dispersion in Fractured Me-
dia. Water Resources Research 19:1253-1265.
Shapiro, A.M. 1987. Transport Equations for Fractured Po-
rous Media. In: Advances in Transport Phenomena in
Porous Media, J. Bear and M.Y. Corapcioglu (eds.),
NATO Advanced Study Institutes Series E, Vol. 128,
Martinus Nijhoff Publishers, Dordrecht, The Netherlands,
pp. 407-471.
Shapiro, A.M. 1989. Interpretation of Oscillatory Water-Level
Responses in Observation Wells During Aquifer Tests in
Fracture Rock. Water Resources Research 25(10>2129-
2138.
Shimo, M. and J.C.S. Long. 1987. A Numerical Study of
Transport Parameters in Fracture Networks. In: Flow and
Transport through Unsaturated Fractured Rock, D.D.
Evans and T.J. Nicholson (eds.), AGU Monograph 42,
American Geophysical Union, Washington, DC, pp. 121-
131.
Smith, L. and F.W. Schwartz. 1984. An Analysis of Fracture
Geometry on Mass Transport in Fractured Media. Water
Resources Research 20:1241-1252.
Smith, L., C.W. Mase, and F.W. Schwartz. 1985. A Stochastic
Model for Transport in Networks of Planar Fractures. In:
Greco 35 Hydrogeologie, Ministere de la Recherche et la
Technologies Centre Nationale de la Recherche
Scientifique, Paris.
Streltsova, T.D. 1988. Well Testing in Heterogeneous Forma-
tions. John Wiley & Sons, New York, 413 pp.
van der Kamp, G. 1976. Determining Aquifer Transmisivity
by Means of Well Response Tests: The Underdampened
Case. Water Resources Research 12(l):71-77.
Ward, D.S., D.C. Skipp, D.A. Giffm, andM.D. Barcelo. 1989.
Dual-Porosity and Discrete Fracture Simulation of Ground
Water Flow in West-Central Florida. In: NWWA Confer-
ence on Solving Ground Water Problems with Models
(Indianapolis, IN), National Water Well Association,
Dublin, OH, pp. 385-408.
Witherspoon, P.A., J.C.S. Long, E.L. Majer, and L.R. Myer.
1987. A New Seismic Hydraulic Approach to Modeling
Flow in Fractured Rocks. In: Proceedings, NWWA/
IGWMC Conference on Solving Ground-Water Prob-
lems with Models (Denver, CO), National Water Well
Association, Dublin, OH, pp. 793-826.
Young, S.C. and W.R. Waldrop. 1989. An Electromagnetic
Borehole Flowmeter for Measuring Hydraulic Conduc-
tivity Variability. In: Proc. Conf. on New Field Tech-
niques for Quantifying the Physical and Chemical
Properties of Heterogeneous Aquifers (Dallas, TX), Na-
tional Water Well Association, Dublin, OH, pp. 463-474.
82
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Chapter 7
Geochemical Characterization of the Subsurface: Basic Analytical and Statistical
Concepts
J. Russell Boulding and Michael J. Barcelona
This chapter presents basic analytical and statistical con-
cepts related to the measurement and interpretation of geo-
chemical data on the natural and contaminated subsurface
environment. Many expensive geochemical investigations suf-
fer because analytical and statistical variability may have been
ignored or not fully appreciated in the sample design and
collection phase. Consequently, these analytical and statistical
concepts are covered here before the chapters on the subsur-
face geochemical variability (Chapter 8), and the best meth-
ods for sampling the subsurface to characterize this variability
(Chapter 9). In the normal sequence of events, laboratory
amlysis and data interpretation come after sample collection.
However, because they should be carefully considered in the
design of geochemical investigations they are presented here
first.
7.1 Data Measurement and Reliability
7.1.1 Deterministic versus Random
Geochemical Data
Observation or measurement of physical phenomena can
be broadly classified as either deterministic or nondeterministic.
Deterministic data can be described by an explicit mathemati-
cal relationship. Nondeterministic or random data, must be
described in terms of probability statements and statistical
averages rather than by the use of explicit equations. Figure 7-
1 summarizes a classification scheme for deterministic and
random data from Bendat and Piersol (1986). The classifica-
tion of physical data as deterministic or nondeterministic is
not always clear-cut in the real world. In fact, most geochemi-
cal data probably fail in a gray area between the two types of
data. For example, the total dissolved solids in an aquifer is a
function of the chemical composition of the aquifer solids and
residence time of the flowing ground water. Consequently, the
distribution of sample values over space and time will not be
completely random. On the other hand, the factors that deter-
mine the precise value of a given sample are sufficiently
complex and variable that the distribution often cannot be
predicted by an explicit mathematical equation.
The transient, nonperiodic data box in Figure 7-la is a
residual category that includes all data not included in the
other boxes, This nonperiodic characteristic of geochemical
data allows modeling of the distribution of geochemical spe-
cies using thermodynamic principles. Essentially all geo-
chemical modeling of the subsurface is done deterministi-
tally. The difficulty in accurately modeling the geochemistry
of the subsurface can, however, be attributed to large random
elements (see Figure 7-lb). Depending on the geochemical
parameter, and the time frame of sampling, data may be
stationary, where characteristics of the population being
sampled do not vary over time, or nonstationary, where the
random process varies with time. Typically, geochemical
subsurface data involving contamination are nonstationary,
but are not fully random (i.e., the value of one sample may
show some correlation with the value of an adjacent sample).
This creates special considerations in statistical analysis that
are discussed in Section 7.3. Subsurface physical parameters
such as hydraulic conductivity, porosity, and soil particle size
distribution do not typically change with time, at least not on
a time scale of human concern. These parameters, however,
are not fully random.
7.1.2 Data Representativeness
In measuring environmental parameters, there is no "true"
value, but rather a distribution of values. A representative unit
or sample is one selected for measurement from a target
population so that it, in combination with other representative
samples, will give an accurate picture of the phenomena being
studied (Gilbert, 1987). Failure to take samples from locations
and to use methods that yield samples that are "representa-
tive" of a site will result in the collection, at some expense, of
analytical data that may be worthless. Representativeness
determines whether accurate analysis of the samples will yield
results that are close to actual conditions. Quality assurancd
quality control systems (QA/QC) in the laboratory or field
may be useless if even greater emphasis isn't placed on QA/
QC in selecting locations and procedures for sampling.
Thorough site characterization of soils, hydrology, and
geology, as described in the previous chapters, is an essential
prerequisite to geochemical sampling. This information pro-
vides the basis for developing sampling strategies that will
provide some assurance that geochemical samples accurately
reflect what is happening in the field. Sample representative-
ness is essentially knowledge-based. For example, sampling
locations selected by someone with a rudimentary under-
standing of sampling theory may yield less accurate results
83
-------
Deterministic
Periodic
Nonperiodic
Complex
periodic
Almost-
periodic
Transient
(b)
Special
classifications
of
nonstationarity
Figure 7-1.
Classifications of (a) deterministic and (b) random
data (from Bendat and Piersol, 1986).
than locations chosen by an individual thoroughly grounded
in this theory. At the same time, sampling locations selected
without careful site characterization will yield less representa-
tive samples than locations selected with thorough site charac-
terization even with equally sophisticated application of
sampling theory. See Section 9.1 for general considerations in
designing sampling plans.
In contamination investigations, obtaining samples that
can be considered representative for assessing one or more
particular kinds of environmental exposure is a primary objec-
tive. This requires selecting not only the right place, but the
right type of sample (see discussions of analyte selection in
Sections 9.2.1 and 9.3.1).
7.1.3 Measurement Bias, Precision, and
Accuracy
A measured value that is close to the estimate of the true
average value is an unbiased or accurate value. This average
or mean can only be estimated by a number of repeat determi-
nations. Biased measurements will consistently under- or
overestimate the true values in sampled population units.
Precision is a measure of how closely individual measure-
ments agree and is influenced principally by random measure-
ment uncertainties. Both bias and precision influence accuracy
as illustrated in Figure 7-2. The center of each target in the
figure represents that true value. Both low bias and high
precision are required for high accuracy.
Accuracy is largely technologically based. In other words,
accuracy can be improved by better drilling and monitoring
well installation procedures and better sampling devices and
procedures. Pennine (1988) has suggested that "there is no
such thing as a representative ground water sample" because
of geochemical biases inherent in well installation, purging,
and sample collection. However, a good understanding of
both potential sources of error (see next section) and the way
alternative sampling methods may bias results (see Section
9.3) minimizes sample disturbance. The final evaluation of
the results should be done with full consideration of the
unavoidable disturbances involved in subsurface investiga-
tions.
7.1.4 Sources of Error
Random error results from slight differences in the execu-
tion of the same sampling procedure. Systematic error results
from procedures that alter the properties of the sample. Ran-
dom error is unavoidable, but must be evaluated to determine
its effect on accuracy. For example, Figure 7-2b shows data
with no systematic bias, but accuracy is low because random
error is high. Systematic errors can be minimized by careful
selection and conduct of sampling techniques.
Figure 7-3 illustrates five possible sources of error in
ground-water sampling: (1) site selection, (2) sampling,
(c) (d)
Figure 7-2. Shots on a target analogy for illustrating influence
of bias and precision on accuracy (after Jessen,
1978).
(a): high bias + low precision= low accuracy; (b):
low bias + low precision = low accuracy; (c): high
bias + high precision = low accuracy; (d) low bias
+ high precision = high accuracy.
-------
Site Selection
Sampling
S!
s-'- !*.«*
Measurement
Methods
Data
Handling
S!
d
Reference
Samples
777US the overall variance = S*=Sljl+S*+S1m+S?+Sd
Figure 7-3. Sources of error involved in ground-water monitoring programs contributing to total variance (from Barcelona et al.,
1983).
(3) measurement methods, (4) reference samples for calibra-
tion, and (5) data handling. Both random and systematic
errors may be involved in each stage. Errors at each stage are
cumulative, but are not of equal significance or magnitude.
Total variance in geochemical data results from the combina-
tion of natural geochemical variability and the cumulative
error. The percentage of variance attributable to natural vari-
ability may often be greater than either field or laboratory
error. Natural variance cannot be reduced; however, variance
resulting from field and laboratory error can be reduced so
that the actual variance closely approximates the natural vari-
ance.
Table 7-1 shows estimates of the relative contribution of
natural variability, field error, and laboratory error to total
variance at three sites of ground-water investigations. For
most chemical constituents, at the three sites, natural variabil-
ity accounted for more than 90 percent of the variance. For
most inorganic constituents where field and laboratory error
could be estimated, field error contributed a larger percentage
of total variance. Table 7-1 also shows that organic contami-
nant indicators (TOC and TOX) showed typically much higher
percentages of variance due to field and laboratory error than
did the inorganic indicators. Both field sampling and labora-
tory analyses were subject to strict QA/QC procedures at the
sites shown in Table 7-1, so variance due to field and labora-
tory error during routine ground-water investigations will
commonly be greater than shown in the table.
Field Error. Figure 7-4 identifies specific possible sources
of error at various steps in ground-water sampling. The largest
sources of error are unrepresentative sample locations (hence
the importance of hydrogeologic site characterization prior to
geochemical sampling design) and disturbances caused by
drilling and well construction. Sample collection is the next
largest source of error. Major sources of systematic sampling
error include (1) well construction and screen design prevent-
ing representative samples, and (2) improper purging. All of
these large sources of systematic error are related to the
hydrology of the site over which there is often little QA/QC.
Table 7-2 lists potential contributions of sampling meth-
ods and materials to error in ground-water chemical results.
This table shows that well purging procedures can result in
large variations in pH, TOC, Fe(II), and VOCs (also see
Table 9-11 for variations in other constituents). Table 7-2
shows well casing to be the next largest source of error,
followed by sampling mechanisms and grouting/sealing.
Poorly grouted or cemented wells can greatly alter the pH of
water (as much as pH 12). Sampling tubing can result in
errors in VOC measurement. Sections 9.3.3 (Purging) and
9.3.4 (Well Construction and Sampling Devices) discuss
selection criteria for minimizing error from these sources.
Other possible sources of systematic error in sampling
include (1) changing sampling procedures, (2) changing
sampling personnel without a strictly defined sampling pro-
tocol, and (3) failure to document unavoidable deviations
from the sampling protocols, such as no water in the well.
Another source of water quality error is mixing from mul-
tiple aquifers. Mixing is most common with public water
supply wells that penetrate several hydrological unconnected
aquifers. Improper sealing of ground-water monitor wells
also may bias results by mixing water from distinct subsur-
face formations.
Analytical Error. Figure 7-5 identifies possible sources
of error during water sample analysis. Analysis, including
measurement methods and reference samples, is typically
subject to the most stringent QA/QC procedures, and conse-
quently analytical errors tend to be relatively minor compo-
nents of total error (see Table 7-1). Failure to analyze blanks,
standards, and samples by exactly the same procedures may
result in either a biased blank correction or a biased calibra-
tion (Kirchmer, 1983). Porter (1986) examined in detail the
sources of random analytical error for measurement near the
limit of detection and how to incorporate this observation
error into data analysis procedures. Sources of analytical
error are discussed further in Section 7.3.
Einarson and Pei (1988) and Rice et al, (1988), in
separate studies of laboratory performance, concluded that
the reliability of laboratory analyses should not be taken for
granted. Both studies also concluded that the cost of analysis
did not necessarily correlate with analytical accuracy. The
most expensive of the 10 laboratories evaluated by Einarson
85
-------
Table 7-1. Percentage of Variance Attributable to Laboratory Error, Field Error, and Natural Variability by Chemical and Site
Type of
parameter
lab
Sand Ridge
Beardstown (upgradient)
Beardstown (downgradient)
field
nat
lab
field
nat
lab
field
nat
Water quality
NO,
SO-
SO
o-PO;
T-PO;
C!
Ca
Mg
Na
K
Geochemical
NH,
NO,
S =
Fe"
Fe
0.0
0.0
0.0
1.2
0.0
7.2
0.0
0.0
0.0
0.0
0.0
NA
NA
NA
0.0
0.0
Contaminant
indicator
roc
TOX
15.4
0.0
00.0
0.0
NA
1.2
NA
NA
45.7
20.0
NA
NA
0.0
NA
NA
NA
NA
NA
lab + field"
100.0
100.0
100.0
97.6
100.0
32.S
54.3
80.0
100.0
100.0
100.0
NA
NA
NA
100.0
100.0
84.6
100.0
0.1
0.2
0.0
0.0
2.8
0.0
0.0
0.0
0.0
33.9
0.0
0.1
NA
0.0
0.0
0.0
29.9
12.5
NA'
NA
20.0
0.0
NA
3.3
2.3
2.2
0.3
NA
0.0
NA
NA
0.1
0.0
40.1
lab + field
99.9
99.8
80.0
100.0
97.8
96.7
97.7
97.8
99.7
66.1
100.0
99.9
NA
99.9
100.0
59.9
70.1
87.5
0.2
1.4
0.0
0.0
0.9
0.0
0.0
0.0
0.0
87.1
0.0
0.3
NA
0.0
0.0
0.0
40.6
24.6
NA
0.1
6.8
0.0
NA
17.2
3.6
2.8
7.1
NA
0.0
NA
NA
5.9
NA
73.6
lab + field
99.8
98.6
93.2
100.0
99.1
82.8
96.4
97.2
92.9
12.9
100.0
99.7
NA
94.1
100.0
26.4
59.5
75.4
"NA indicated that the number of observations on which the estimated variance was based was less than 5, or the estimated variance was
negative.
b True field spiked standards not available for these costituents demanding combined estimates of laboratory and field variability.
Source: Barcelona et al. 1989
Step
In-Situ Condition
Establishing a Sampling Point
i
Field Measurements
Sample Collection
Sample Delivery/Transfer
4
Field Blanks, Standards
I
Field Determinations
Preservation/Storage
•A-
Transportation
Sources of Error
Improper well construction/placement;
inappropriate materials selection
Instrument malfunction; operator error
Sampling mechanism bias; operator error
Sampling mechanism bias; sample exposure,
degassing, oxygenation; field conditions
Operator error; matrix interferences
instrument malfunction; operator error;
field conditions
Figure 7-4.
Mattrix interferences; handling/labeling errors
Delay; sample loss
Steps in ground-water sampling and sources of error (from Barcelona et al., 1985).
86
-------
Table 7-2. Potential Contributions of Sampling Methods and Materials to Error" in Ground-Water Chemicail Results
Sampling method/
material
Range of
concentration
Drilling muds
Grouts, seals
Well purging
pH
5-9
—
+, 4 to 5
units
(cement)
±, 0.1 to
5 units
roc
(mg C/L)
0.5-25
+.300%
±,500%
Fe(ll)
(mg/L)
0.01-10
—
-," 500%
cement
-," 1000%
voc
(W/L)
0.15-8000
...
±, 10 to 1000%"
Well casing
Sampling
mechanism
Sampling
tubing
References
gas lift +,
0. 1 to 3
units
1,5,7
±,200%
bailer +,
150%
1,4
+, 1000%
iron,
galvanized
steel
gas lift-,11
500%
1,2,5,7
±,200%'
suction -,*
1 to 15%"
-10 to 75%"
1,3,6
'Bias values exceeding >± 100% denoted as gross errors ( + or -): other values expressed as percent of reported mean.
"No data available on the type and extent of error for this parameter.
'Concentration range 0.5-15 pg/L (from Barcelona and Helfrich, 1984).
"Concentration range 80-8000 ug/L (from Barcelona et al., 7984; Ho, 1983).
1 Barcelona and Helfrich (1984)
2 Barcelona et a/. (1983)
3 Barcelona et al. (1984)
4 Barcelona et al. (1988)
5 Gibbet al. (1981)
6 Ho (1983)
7 Schuller et al. (1981)
Source: Adapted from Barcelona et al. (1988)
and Pel (1988) tied for the worst ranking, while the four least
expensive laboratories included the top ranked and other
bottom ranked laboratory. Both studies describe criteria and
procedures for choosing laboratories that will provide good
analytical results. Section 7.3 discusses analytical and QA/QC
concepts further.
Date Handling Error. There is probably no large body of
scientific records free from human or machine errors. Faulty
recording of observations in field or laboratory notebooks or
incorrect coding for computer analysis are examples of data
handling errors. Misrecorded values that are much larger or
smaller than the range of the actual population are called
outliers and may distort the results of statistical analysis.
Statistical techniques are available for analyzing such data
sets (Gilbert, 1987), but prevention of data handling error is
always better than a cure. Censoring of analytical measure-
ments below the limit of detection (see Section 7.4.1) is
another serious error introduced by data handling.
Webster (1977) suggests some of the following methods
to reduce data handling errors: (1) write neatly, forming
characters well; (2) distinguish ambiguous digits and letters
by a firm convention; (3) restrict the digit 0 to mean zero and
use other notations for "missing" or "inapplicable"; (4) elimi-
nate or minimize transcription of field notes (5) record data
on forms designed for the purpose of the investigation with
clear headings and ample space; and (6) double-check any
transcribed data against the original
7.2 Analytical and QA/QC Concepts
Quality assurance and quality control are accomplished
by (1) selecting the best methods for the program purpose, (2)
clearly defining protocols or procedures to be followed, and
(3) carefully documenting adherence or departures from the
protocols. Figure 7-6 shows the relationship of program pur-
pose and protocols to the scientific method. Both field sam-
pling and laboratory analyses require protocols for good QA/
QC. Campbell and Mabey (1985) have summarized key ele-
87
-------
Step
Sources of Error
Samples, from Storage
Field Blanks and Standards
4
Subsampling
4
Procedural Standards
4
Analytical Separation
Analysis
I
Reference Standards
4
Calculations
Results
"Aged" samples; loss ofanalyties;
contamination
Sample aging/contamination in lab; cross-
contamination; mishandling/labeling
"Aged" standards; analyst error
Matrix interferences; inappropriate/
invalid methodology; instrumental
malfunction/analyst error
Matrix interferences; inappropriate/
invalid methodology; instrumental
malfunction/analyst error
"Aged" standards
Transcription/machine errors; sample loss in
tracking system; improper extrapolation/
interpolation; over-reporting/
under-reporting errors
Figure 7-5. Steps in water sample analysis and sources of error (from Barcelona et al., 1985).
Hypothesis
Program
Purpose
Formulate -t
Questions and
Design
\
Observation
Sample
- Sampling-*'
Protocol
1
Procedures
\
Techniques
\
Methods
Analyze
Analytical -
Protocol
t
procedures
I
Techniques
t
Methods
Interpretation
Interpret
»- Results
Figure 7-6. Relationship of program purpose and protocols to
the scientific method (from Barcelona, 1988).
ments of data evaluation systems applicable to both field and
laboratory measurements. Provost and Elder (1985) have pro-
vided guidance for choosing cost-effective QA/QC programs
for chemical laboratories. Evans (1986) reviews data quality
objectives for remedial site investigations, and Starks and
Flatman (1991) discuss the use of industrial quality control
methods as a model for evaluating RCRA ground-water moni-
toring decision procedures.
7.2.1 Instrumentation and Analytical Methods
A bewildering array of methods are available for analyz-
ing geochemical constituents. Table 7-3 lists the major signals
and analytical methods based on signal measurement. Most
methods used for geochemical analysis involve either emis-
sion or adsorption of radiation. The fine points of instrumenta-
tion and analysis are the province of the analytical chemist,
but the field scientist can benefit from a general understand-
ing. Skoog (1985) and Willard et al. (1988) are two good
general references on this topic. Analytical techniques for
specific constituents of geochemical interest may be specified
by regulation or, if not so specified, determined by the instru-
mentation that is most readily available. Table 7-4 lists seven
major sources of information describing analytical techniques
for specific chemical constituents.
7.2.2 Limit of Detection
Ground-water detection monitoring commonly involves
measurement of contaminants that are either at or below the
detection limit of analytical procedures. The statistical con-
cept of detection limit includes accurately reporting and ana-
Ivzing data including measurement near or below the detection
limit (McNichols and Davis, 1988).
-------
Table 7-3. Major Analytical Signals and Methods
Signal Analytical Methods Based on
Measurement of Signal
Emission spectroscopy (X-ray, UV, visible
electron auger); fluorescence and
phosphorescence spectroscopy (X-ray, UV,
visible); radiochemistry
Spectrophotometry (X-ray, UV, visible, IR);
photoacoustic spectroscopy; nuclear
magnetic resonance and electron spin
resonance spectroscopy
Turbidimetry; nephelometry; Raman
Spactroscopy
Refractometry; interferometry
Emission of
radiation
Absorption of
radiation
Scattering of
radiation
Refraction of
radiation
Rotation of
radiation
Electrical
potential
Electrical
current
Mass-to-charge
ratio
Rate of reaction
Thermal
properties
Mass
Volume
Polarimetry; optical rotatory dispersion;
circular dichroism
Potentiometry; chronopotentiometry
Polarography; amperometty; coulometty
Mass Spectrometry
Kinetic methods
Thermal conductivity and enthalpy methods
Gravimetric analysis
Volumetric analysis
"Source: Skoog (1985)
Figure 7-7 and Table 7-5 illustrate the definitions of limit
of detection and regions of analyte measurement recom-
mended by the Subcommittee on Environmental Analytical
Chemistry of the American Chemical Society's (ACS) Com-
mittee on Environmental Improvement (1980). The zero analyte
signal for measuring the limit of detection comes from the
field blank (see Section 7.2.3). If the actual field blank mea-
surement gives a positive signal, this means that analytical
measurements on other samples with a lower signal will be
recorded as a negative concentration. For example, a low
concentration standard (typically 1 part per billion (ppb) for
organic constituents) is made in the laboratory for the con-
taminant of interest. The standard deviation for analytical
measurement of the 1 ppb standard is commonly plus or
minus 100 percent or 1 ug/L. The detection limit for a
contaminated sample is defined as three standard deviations
(3 ug/L) above the mean for the standard, or six standard
deviations above the zero point defined by the field blank (see
Figure 7-7). The limit of detection should be defined every
day of analysis. The detection limit is probably the most
important kind of laboratory quality assurance data and should
be reported with the analytical results for each constituent.
Table 7-5 lists the regions of analyte measurement. Fol-
lowing the above example, signals below three standard de-
viations are considered below the limit of detection. The
region of detection is between 3 and 10 standard deviations (5
standard deviations by some rules) and is where the constitu-
ent can be said to be present but the precise concentration
cannot be stated with certainty. Analyte signals above the
limit of quantification (plus 10 standard deviations) can be
interpreted quantitatively.
The above-described definition reaffirms the model for
limit of detection calculations adopted by the International
Union of Pure and Applied Chemistry (IUPAC) in 1975
(IUPAC, 1978). However, considerable confusion still sur-
rounds the definition of the limit of detection. This is because
(1) acceptance of the above definition by the general analyti-
cal community has been slow, and (2) different statistical
approaches to calculating limits of detection for constituents
can easily vary by an order of magnitude (Long and
Winefordner, 1983). This is particularly true for chemical
constituents at the ppb level.
A major problem with failure to understand the statistical
nature of the limit of detection is negative censoring of data.
Negative censoring involves reporting analyte concentrations
that are below the limit of detection as zero, "less than"
values, or "not detected." Since 1983 the American Society
for Testing and Materials (ASTM) has recommended that data
should not be routinely censored by laboratories (ASTM,
1983). Nevertheless, censoring of water quality analytical
data remains a problem (Porter et al., 1988). Section 7.4.1
examines this issue further.
Laboratories should be asked to provide uncensored data
on all water samples with measurements near or below the
limit of detection. Measurement data should not be discarded
unless the lack of statistical control in the measurement pro-
cess is clearly demonstrated. The general public, and even the
uninformed scientist, may find the concept of a negative
concentration difficult to understand, so it is prudent to report
less than zero values as "trace." Remediation decisions, how-
ever, should be based on concentrations at or above the limit
of quantification, not the limit of detection.
The limit of detection is both a site- (as a result of the
field blank) and instrument/operator-specific value. Conse-
quently, the precision and accuracy for low standards must be
reported on the analytical report forms. The instrument
manufacturer's definition of detection is based normally on
carefully controlled conditions (e.g., distilled water solutions)
that may not be achievable in routine analyses of complex
samples. Consequently, actual limits of detection in contami-
nated ground water are often higher.
7.2.3 Types of Samples
Field scientists tend to consider QA/QC requirements and
procedures to be primarily the responsibility of the laboratory.
However, QA/QC procedures are equally, if not more impor-
tant in the field. Chapter 8 examines methods to minimize
error in selecting sample location and collecting samples.
Field personnel also should be familiar with the different
types of samples that may be taken, and their importance for
interpreting the analytical results.
-------
Table 7-4. Major Compilations of Analytical Procedures for Constituents of Geochemical interest
Reference Description
American Public Health
Association (1990)
ASTM, annual
Fresenius et al. (1988)
Klute (1986), Page et al.
(1982)
Kopp and McKee (1983)
Longbottom and
Lichtenberg (1982)
Mueller et al. (1991)
Noblett and Burke (1990)
Radian Corporation
(1988)
Rainwater and
Thatcher (1960)
Smith (1991)
Thompson et al. (1989)
U.S. EPA (1988)
U.S. Geological Survey
Techniques of Water-
Resource Investga-
tions
Westerman (1990)
Comprehensive compilation of analytical methods for measurement of metals, inorganic nonmetallic, and
organic constituents in water samples.
Published annually by the American Society for Testing and Materials, Water and Environmental Technology
Volumes 11.01 and 11.02 cover analytic methods for water.
A guide to physico-chemical, chemicaland microbiological analysis of water and qualify assurance procedures.
Part 1 (Klute, 1986) contains 50 chapters covering a range of physical and mineralogical methods and Part 2
(Page et al., 1982) contains 54 chapters covering methods for analyzing chemical and microbiological
properties of soils.
This third edition contains the chemical analytical procedures used in U.S. EPA laboratories for examining
ground and surface water, domestic and industrial waste effluents and treatment process samples.
Describes tests for 15 groups of organic chemicals and includes an appendix defining procedures for
determining the detection limit of an analytic method. The test procedures in this manual are cited in
Table 1C (organic chemical parameters) and ID (pesticide parameters) in 40 CFR 136.3(a).
Compilation of summary information on more than 150 EPA-approved, and a total of 650, sampling and
analysis methods for industrial chemicals, pesticides, elements, and water quality parameters. Associated
data base is available on diskette.
Handbook on flue gas desulfurization (FGD) chemistry and analytical methods. Volume 1 (Noblett and
Burke, 1980) covers sampling, measurement, laboratory, and process performance guidelines. Volume 2
(Radian Corporation, 1988) presents 54 physical-testing and chemical-analysis methods for FGD reagents,
slurries, and solids.
Describes types of methods, choice of analytical methods for water samples, and specific analytical
procedures for over 40 inorganic water parameters.
Edited volume with 14 chapters on instrumental techniques for soil analysis.
Contains summary description of methods for elemental analysis, analysis of anionic species, inorganic and
organic carbon, redox sensitive species and other chemical parameters along with recommendations for
methods best suited for obtaining data for hydrochemical modeling.
Guide for selection of instrumental methods for field screening of inorganic and organic contaminants. Covers
26 specific field screening methods. Also available as a computerized information retrieval system.
USGS's TWI series includes manuals describing procedures for planning and conducting specialized work in
water-resources investigations. Wood (1976) covers field analysis of unstable constituents; Skougstad et al.
(1979) cover methods for analyzing inorganic constituents in water and fiuvial sediment; Bamett and Mallory
(1971) describe determination of minor elements in water by emission spectroscopy: Wershaw et al (1987)
cover methods for determination of organic substances in water and fiuvial sediments
(revision of Goeriitz et al., 1972).
Edited volume on methods for analysis of soil and plants focussing on use for assessing nutritional
requirements of crops, efficient fertilizer use, saline-sodic conditions, and toxicity of metals.
A field blank is a sample of distilled or deionized water
taken from the laboratory out into the field, poured into a
sampling vial at the site, closed, and returned as if it were a
sample. The level of contamination of the field blank is the
zero analyte signal for determining the limit of detection,
A rinse or cleaning blank is a sample of the final rinse of
a sampling mechanism before it is put in a new well. This type
of sample is used to evaluate whether a sample may have been
contaminated from material taken in the previous sample.
Field samples are those samples that are taken in the field
as "representative" of conditions at the site and analyzed in
the laboratory for constituents of interest. If sampling points
or locations are unrepresentative, or biased sampling proce-
dures are used, no amount of care in QA/QC in subsequent
stages will salvage an accurate picture of actual field condi-
tions.
Duplicate samples are collected and not analyzed unless
it is later determined that they contain additional useful infor-
mation. Soil samples are commonly duplicated.
Replicate samples are subsamples of the same sample
that are labeled separately to estimate the precision of labora-
tory analytical results.
Split samples are field samples that are split between two
storage vessels or cut in half in the field. One subsample may
be analyzed by one laboratory and the other subsample may
be archived or given to another laboratory,
90
-------
• Analyte Signal (Sx)'
Table 7-5. Regions of Analyte Measurement
Analyte Signal Recommended Inference
(standard deviations in \ng/L)
Zero '
Analyte
Not Detected
Zero Sb Sb -
LC
•„ £
Region of
Detection
-------
Geostatistical techniques have three main applications for
characterization of subsurface variability: (1) they can assist
in reducing spatial sampling intensity, and hence reduce sam-
pling and analytical costs; (2) they can be used to differentiate
sample data that are autccorrelated or noncorrelated, elucidat-
ing trends for selecting the appropriate statistical analysis of
sampling analytical results; and (3) they can be used to
interpolate values at locations where measurements have not
been made. The last application is done by kriging, a weighted
moving-averaging technique, that in most situations will pro
vide the most accurate way of contouring data on physical and
geochemical parameters. Furthermore, a kriging standard de-
viation map that provides a clear indication of the reliability of
contours can be readily created from kriged contour data.
One of the first steps in geostatistical analysis is to
calculate the nonsampling variance (gamma) of samples at
different distance spacings. Gamma is a statistical measure of
the difference between sample values. For example, if samples
were taken from a 50-m grid, gamma would be calculated for
the samples spaced at 50 m, 100 m, 150 m, 200 m, and so on.
Next, a semivariogram is plotted on a XY plot, where X is
distance and Y is the nonsampling variance. Figure 7-8 shows
an "ideal" semivariogram. Samples within a certain range of
influence, also called the range of correlation (distance a in
Figure 7-8), show an approximately linear correlation (are
autocorrelated). At some spacing distance, if there is no trend
in the data, a sill (C on Figure 7-8) marks a plateau that limits
the range of correlation. The nonsampling variance between
samples will equal C as long as the distance is greater than a.
From a sampling perspective, samples spaced closer than
distance a in Figure 7-8 will yield redundant, correlated data,
which results in both unnecessary expense and complications
in statistical analysis. The minimum distance at which samples
Figure 7-8. The '"ideal" shape for a semivariogram-spherical
model (from Clark, 1979).
are independent (distance a in Figure 7-8) is the optimum
sampling distance.
Figure 7-9 shows a semivariogram of lead values in soil
sampled by Flatman (1986) on a systematic 750-ft grid. The
diagram shows that samples for lead that are closer to each
other than about 1,200 ft are correlated. In other words, the
same information could be obtained by cutting the number of
samples almost in half. Figure 7-10 shows a kriged contour
map of lead concentrations in the vicinity of the smelter, and
Figure 7-11 shows contours of the standard deviations of the
lead concentrations.
Table 7-6 summarizes ranges of influence (in meters) that
have been estimated for a variety of soil physical and chemi-
cal parameters. Direct comparisons between different studies
are difficult, however, because definitions and the methodolo-
gies for determining the range vary somewhat. Commonly,
however, the range is scale-dependent, i.e., as the sample area
increases, the range increases. For example, at the same site
Gajem et al. (1981) found ranges of 1.5,21, and 260 m for pH
values of 100-member transects spaced at 0.2, 2, and 20 m.
Semivariograms may exhibit a variety of correlation struc-
tures other than the one shown in Figure 7-8, and correct
interpretation requires an understanding of the various models
that are available for describing semivariogram plots. When
data are not normally distributed, such as when a spatial trend
is present, estimating the correlation structure is difficult. In
these cases, some of the techniques for transforming log-
normal data for conventional statistical analysis can be used
(Gilbert, 1987).
Most basic texts on geostatistics are still oriented towards
mining. Clark (1979) provides a good introduction to
geostatistics and kriging, while more comprehensive treat-
ments (all oriented toward mining) can be found in the follow-
ing sources: David (1977), Isaaks and Srivastava (1989),
Matheron (1971), and Journal and Huijbregts (1978). Olea
(1974, 1975) provide a good introduction to the use of geosta-
tistics in contour mapping of data. Gilbert and Simpson (1985)
provide a good review of potentials and problems with using
kriging for estimating spatial pattern of contaminants
Trangmar et al. (1985) and Warrick et al. (1986) re-
viewed specific geostatistical methods applied to spatial stud-
ies of soil properties. Use of geostatistics in sampling for soil
contaminants is discussed by Flatman (1984), Flatman and
Yfantis (1984), and Flatman (1986). Delhomme (1978, 1979)
reviewed the use of geostatistics in the characterization of
ground-water variability, and Hughes and Lettenmaier (1981)
and Sophocleous et al. (1982) discuss applications for ground-
water monitoring network design.
7.4 Interpretation of Geochemical and Water
Chemistry Data
Table 7-7 indexes some sources of information on (1)
basic statistical approaches to data analysis, (2) methods for
analysis of soil data, and (3) methods for analysis of water
quality data. The general references on soil and water chemis-
try listed in Table 7-1 provide a framework for interpreting
92
-------
It
Tickmark = 167m (500 feet)
Lag (the Distance between Sample Locations)
Figure 7-9. A semivariogram of lead samples taken systematically on a 230-m (750-foot) grid (from Flatman, 1986).
background geochemistry. Hem (1985) is an especially good
source for the interpretation of water quality data.
Gilbert (1987) presented probably the best systematic
treatment of statistical methods for environmental pollution
monitoring. Bury (1975) provides a comprehensive treatment
of basic statistical concepts and models oriented toward the
applied scientist. Hollander and Wolfe (1973), Lehmann and
D'Abrera (1975), and Seigel (1956) offer more in-depth dis-
cussion of nonparrametric statistical methods. Bury (1975)
presents a table that is a useful guide for finding the appropri-
ate nonparametric procedure for particular topics or problems.
Chatfield (1984) is a good source on techniques for analysis of
time series.
7.4.1 Analysis of Censored Data
Table 7-8 illustrates the effect of two types of censoring
of analytical results near and below the limit of detection.
Data reported as less than the limit of detection are heavily
censored and yield an average concentration of 3.5 ug/L
since only two values are quantified. Reporting of negative
concentrations as zero is called negative censoring; in Pable
7-8 negative censoring yields an average of 1.2 ug/L. The
uncensored data average 0.5 (ig/L. The averages of the heavily
and negatively censored data would appear to indicate con-
tamination, but the 95 percent confidence interval for the
uncensored data is at best equivocal.
Gilliom et al. (1984) found that any censoring of trace-
level water quality data, even when the censored data were
highly unreliable, reduced the ability to detect trends in the
data. Unfortunately, censored data continues to be routinely
reported by laboratories. The following references contain
discussions of statistical techniques for analyzing censored
data: Gilbert (1987), Gilliom and Helsel (1986), Gilliom et al.
(1984), Helsel and Gilliom (1986), McBean and Rovers (1984)
and Porter et al. (1988).
7.4.2 Contaminant Levels versus Background
Conditions
Numbers on a standard list from an analytical laboratory
arc useful only to the extent that they can be compared to
known or estimated background conditions before contamina-
tion. Using such numbers effectively requires both data on
background conditions and the use of appropriate techniques
to detect statistically significant departures from background
levels. An analytical result from a rinse or cleaning blank
between the limit of detection and the limit of quantification
may indicate that more careful decontamination procedures
should be followed, but does not add to the information on
which to base remediation decisions.
Crustal and natural background abundances of metallic
elements must be considered when evaluating analyses for
inorganic contaminants. See the listing under "background"
for soil chemical parameters and water chemistry in Table 8-
1, which identifies some sources of background data on minor
and trace elements in the United States. For organics, there is
always some background of total inorganic carbon and or-
ganic carbon, which should be determined in some samples to
identify natural background levels. The amount of organic
matter may vary considerably in soil, but dissolved organic
93
-------
7300.0
Contour Map of Lead
Concentrations in PPM
2900.0 4500.0
6100.0
7700.0
carbon in ground water does not vary greatly. There are
definite analytical difficulties in achieving reliable analyses in
the range of 0.1 to 0.5 percent organic carbon in the solid
fraction.
Equilibrium calculations based on thorough chemical
analysis may be useful for interpreting water quality data
(Jenne, 1979; Melchior and Bassett 1990 Summers et al.,
1985). For example, reducing or suboxic conditions, indicated
by low Eh (i.e., measured oxidation-reduction potential), lack
of detectable dissolved oxygen, and presence of ferrous iron,
may indicate conditions favorable for movement of elements
such as manganese, mercury, chromium, and arsenic. Arsenic
(V) under oxidizing conditions may be considered immobile,
but under reducing conditions, arsenic (III) is often the pre-
dicted "stable" species of arsenic and is frequently more toxic
and more mobile than As(V) due to higher volubility (Holm
and Curtis, 1984).
Figure 7-10. Kriged contour map of lead concentration in ppm
around a smelter (from Flatman, 1988).
94
-------
500.0
500.0 2700.0 3700.0 5300.0 6900.0 8500.0 10100.0 11700.0
Kriging Error Map-RSP
Figure 7-11. Kriging standard deviation map for lead concentrations around a smelter (from Flatman, 1986).
95
-------
Table 7-6. Reported Values of Ranges of
Source Parameter
Burgess and Sodium
Webster (1980)
Depth cover loam
Campbell (1978) Sand content
Sand content
SoilpH
Clifton and Log of transmlssMty
Neuman (1982)
Folorunso and Flux of N2 and
Rolston (1984) N,O at surface
Gajem ei al. Sand content
(1981)
SoilpH
Kachonoski Depth and mass
etal. (1985) ofA-horizon
McBrainey and pH
Webster (1981)
Russo and Saturated conductivity
Bresler (1981a,
1981b) Saturated water content
Sorptivity
Wetting front
Sisson and Steady-state infiltration
Wierenga (1981)
van Kuilenberg Moisture supply
etal. (1982) capacity
Vauclin et al. Surface soil temperature
(1982)
Vauclin et al. Sand content
(19S3)
pF2.5
Vieira et al. Steady-state infiltration
(1981)
Wollum and Log of most probable
Cassel (1984) number of Rhizobium
Correlation of Soil Physical and Chemical Properties
Range or Site
Scale (m)
61
100
30
40
Random
9600
<1
>5
7.5
21
260
<2
20
34
14
28
37
39
16-30
0.13
600
8-21
35
25
50
Approx. 50 ha, Plas Gogerddan (Gr. Britain), 440 samples, 0-1 5 cm depth
Approx. 18 ha, Hole Farm (Gr. Britain), 450 observations
Lady smith series, me sic Pachic Argiustolls (Kansas), 8x20 grid at 10-m
spacing in B2 horizons
Pawnee series, mesic Aquic Argiudoll (Kansas) (as above)
Pawnee and Ladysmith
Avra Valley (Arizona), about 15x50 km, 148 wells
Yolo loam, Typic Xerorthents (California) 1 00 x 100m area
Pima clay loam, Typic Torrifiuvenis (Arizona), 20-m transect, 20-cm spaces,
50-cm depth.
Pima, as above, 4 transects
Pima, as above but 4 transects, 2-m spacing
Pima, as above, 1 transect, 20-m spacing, 100 points
Mix of Typic Haploborolls and Typic Argiborolls (Saskatchewan)
Surface. Harma Red Rhodoxeralf (Israeli. 30 random sites in 0,8 ha
90-cm depth, as above
90-cm as above
Surface
90-cm, as above
Simulated for above site, 1 to 12.5 h
Sandy clay loam, Typic Torrifluvents (New Mexico), 6.4 x 6.4 m plot,
transect of 125 contiguous 5-cm rings
Cover sand, 30 mapping units, 9 soil types including Haplaquepts,
Humaquepts, and Psammaquents (Netherlands), 2 by 2 km, 1191
borings
Yolo loam clay, Typic Xerorthents (California), 55 x 160 m area
Sandy clay loam (Tunisia), 7x4 grid at 10-m spacing, 20-40 cm depth
Same
Yolo loam, Typic Xerorthents (California), 55 x 160 1 area
Pocalla loamy sand, thermic Arenic Plinthic Paleudults (N. Carolina)
japonicum
Random
0 °, 3-m spacing
0 °, 20-cm spacing
90 °, 3-m spacing
90 °, 20-cm spacing
Yost etal. (1982)
SoilpH
Phosphorus sorbed
at 0.02 mg P/L
Phosphorus sorbed
at 0.2 mg P/L
14,000-
32,000
32,000
58,000
Various transects on Island of Hawaii at 1 to 2 km intervals,
As above
As above
10-15 cm depth
Source: Adapted from Warrick et al. (1986)
96
-------
Table 7-7. Sources of Information
Topic
on Techniques for Anailzing Soil and Water-Quality Data
References
Basic Statistical Approaches
General
Nonparametrics
Time series
Exploratory data
(Median-Polish)
Geostatistics (basic)
Geostatistics (adv.)
Soil Data Analysis
Population properties
Geostatistics
Contaminated soils
Soil Gas Data
Water Quality Data
General
Contaminant detection
Geostatistics
Population properties
Spatial data
Time series data
Bandat and Pierson (1986), Bury (1975), Gilbert (1987), Jessen (1978), tin (1966), Ott (1984)
Hollander and Wolfe (1973), Lehmann (1975), Seigel (1956)
Chatfield (1984)
Tukey (1977), Velleman and Hoaglin (1981), Alhajjar et al (1990)
Clark (1979), Englund and Sparks (1988), Gilbert and Simpson (1965), Journaf (1984),
Olea (1974, 1975), Yates and Yates (1990)
David (1977), Journal and Huijbregts (1978), Isaaks and Srivastava (1989), Matheron (1971)
Butler (1980), Sinclair (1986), Webster (1977)
Sinclair (1986), Trangmar et al. (1985), Warrick et al. (1986). See also Table 7-6
Flatman (1964), Flatman and Yfantis (1984), and Flatman (1986)
See Table 9-5
Beck and van Stratten (1983), Gillham et al (1983), U.S. EPA (1989)
Chapman and El-Shaarawi (1989), Davis and McNichols (1988), Gibbons (1987a,b; 1990),
McBean and Rovers (1990), McNichols and Davis (1988)
Delhomme (1978, 1979), Hughes and Lettenmaier (1981), Samper and Neuman (1985),
Sophocleous et al (1982)
Harris et al. (1987), Montgomery et al. (1987)
Lawrence and Upchurch (1976), McBean et al. (1988)
Close (1989), Harris et al. (1987), McBean et al. (1988), Montgomery et al (1987),
Sgambat and Stedinger (1981), Yevjevich and Harmancioglu (1989)
Table 7-8.
Mean
95% Conf.
Effects of Censoring Analyte Signals at and
Below the Limit of Detection
Sample
1
2
3
4
5
6
7
8
9
10
Heavily
Censored
<3
<3
<3
4
3
<3
<3
<3
<3
<3
Negatively
Censored
2
0
0
4
3
0
r
0
0
2
Uncensored
2
-2
-1
4
3
-3
1
-1
0
2
3.5
0.14-2.26
1.2
1.13
-0.5
-2.13
Source: ASTM (1987)
97
-------
7.5 References
ACS Committee on Environmental Improvement. 1980. Guide-
lines for Data Acquisition and Data Quality Evaluation in
Environmental Chemistry. Analytical Chemistry 52:2242-
2249.
Alhajjar, B.J., G. Chesters, and J.M. Harkin. 1990. Indicators
of Chemical Pollution from Septic Systems. Ground Wa-
ter 28(4):559-568.
American Public Health Association. 1990. Standard Meth-
ods for the Examination of Water and Wastewater, 17th
ed. APHA, Washington, DC.
American Society for Testing and Materials (ASTM). Annual
Books of ASTM Standards. Water and Environmental
Technology, Volumes 11.01 and 11.02 (Water). ASTM,
Philadelphia, PA.
American Society for Testing and Materials (ASTM). 1987.
Standard Practice for Intralaboratory Quality Control Pro-
cedures and a Discussion on Reporting Low-Level Data.
In: Annual Book of ASTM Standards, Vol. 11.01, D4210-
83. ASTM, Philadelphia, PA.
American Society for Testing and Materials (ASTM), Sub-
committee 019.02.1983. Annual Book ASTM Standards,
Volume 11.01, Chapter D, pp. 4210-4283.
Barcelona, M.J. 1988. Overview of the Sampling Process. In:
Principles of Environmental Sampling, L.H. Keith (ed.),
ACS Professional Reference Book, American Chemical
Society, Washington, DC, pp. 1-23.
Barcelona, M.J. and J.A. Helfrich. 1986. Effects of Well
Construction Materials on Ground Water Samples.
Environ. Sci. Technol. 20(11): 1179-1184.
Barcelona, M.J., J.P. Gibb, and R.A. Miller. 1983. A Guide to
the Selection of Materials for Monitoring Well Construc-
tion and Ground-Water Sampling. ISWS Contract Report
327. Illinois State Water Survey, Champaign, IL.
Barcelona, M.J., J.A. Helfrich, E.E. Garske, and J.P. Gibb.
1984. A Laboratory Evaluation of Ground-Water Sam-
pling Mechanisms. Ground Water Monitoring Review
4(2):32-41.
Barcelona, M.J., J.P. Gibb, J.A. Helfrich, and E.E. Garske.
1985. Practical Guide for Ground-Water Sampling. EPA
600/2-85/104 (NTIS PB86-137304). Also published as
ISWS Contract Report 374, Illinois State Water Survey,
Champaign, IL.
Barcelona, M.J., J.A. Helfnch, and E.E. Garske. 1988. Verifi-
cation of Sampling Methods and Selection of Materials
for Ground-Water Contamination Studies. In: Ground-
Water Contamination: Field Methods, A.G. Collins and
A.I. Johnson (eds.), ASTM STP 963, American Society
for Testing and Materials, Philadelphia, PA, pp. 221-231.
Barcelona, M.J., D.P. Lettenmaier, and M.R. Schock. 1989.
Network Design Factors for Assessing Temporal Vari-
ability in Ground-Water Quality. Environmental Moni-
toring and Assessment 12:149-179.
Bardossy, A., I. Bogardi, and L. Duckstein. 1990. Fuzzy
Regression in Hydrology. J. Hydrology 26(7): 1497-1508.
Bamett, P.R. and E.G. Mallory, Jr. 1971. Determination of
Minor Elements in Water by Emission Spectroscopy.
U.S. Geological Survey TWI 5-A2.
Beck, B.F. and G. Van Strateen (eds.). 1983. Uncertainty and
Forecasting of Water Quality. Springer-Verlag, New York.
Bendat, J.S. and A.G. Piersol. 1986. Random Data, Analysis
and Measurement Procedures, 2nd ed. Wiley-Interscience,
New York.
Burgess T.M. and R. Webster. 1980. Optimal Interpolation
and Isarithmic Mapping of Soil Properties I. The
Variogram and Punctual Kriging. J. Soil Sci. 31:315-331.
Bury, K.V. 1975. Statistical Models in Applied Science. John
Wiley & Sons, New York.
Butler, B.E. 1980. Soil Classification for Soil Survey, Chapter
2. Oxford University Press, New York.
Campbell, J.A. and W.R. Mabey. 1985 A Systematic Ap-
proach for Evaluating the Quality of Ground Water Moni-
toring Data. Ground Water Monitoring Review 5(4):58-22.
Campbell, J.B. 1978. Spatial Variability of Sand Content and
pH within Continuous Delineations of Two Mapping
Units. Soil Sci. Soc. Am. J. 42:46044.
Chapman, D.T. and A.H. El-Shaarawi. 1989. Statistical Meth-
ods for the Assessment of Point Source Pollution. Envi-
ronmental Monitoring and Assessment 13 (2/3): 1-467.
[Special issue with 21 papers].
Chatfield, C. 1984. The Analysis of Time Series: Theory and
Practice, 3rd ed. Chapman and Hall, London.
Clark, I. 1979. Practical Geostatistics. Applied Science Pub-
lishers, London.
Clifton, P.M. and S.P. Neuman. 1982. Effects of Kngmg and
Inverse Modeling on Conditional Simulation of the Avra
Valley in Southern Arizona. Water Resources Research
18:1215-1234.
Close, M.E. 1989. Effect of Serial Correction on Ground
Water Quality Sampling Frequency. Water Resources
Bulletin 25(3):507-515.
David, M. 1977. Geostatistical Ore Reserve Estimation.
Elsevier, New York.
-------
Davis, C.B., and R.J. McNichols. 1988. Discussion of "Statis-
tical Prediction Intervals for the Evaluations of Ground-
Water Quality." Ground Water 25(1):90-91.
Delhomme, J.P. 1978. Kriging in the Hydrosciences. Adv.
Water Resources 1:251-266.
Delhomme, J.P. 1979. Spatial Variability and Uncertainty in
Groundwater Flow Parameters: A Geostatistical Approach.
Water Resources Research 15:269-280.
Einarson, J.H. and P.C. Pei. 1988. A Comparison of Labora-
tory Performances. Environ. Sci. Technol. 22:1121-1125.
Englund, E.J. and A.R. Sparks. 1988. Geo-EAS (Geostatistical
Environmental Assessment Software) User's Guide. EPA/
600/4-88/033a (Guide: NTIS PB89-151252, Software:
PB89-151245).
Evans, R.B. 1986. Ground-Water Monitoring Data Quality
Objectives for Remedial Site Investigations. In: Quality
Control in Remedial Site Investigation: Hazardous and
Industrial Solid Waste Testing, Fifth Volume, C.L. Perket
(ed.), ASTM STP 925, American Society for Testing and
Materials, pp. 21-33.
Flatman, G.T. 1984. Using Geostatistics in Assessing Lead
Contamination Near Smelters. In: Environmental Sam-
pling for Hazardous Wastes, G.E. Schweitzer and J.A.
Santolucito (eds.), ACS Symp. Ser. 267, American Chemi-
cal Society, Washington, DC, pp. 43-52.
Flatman, G.T. 1986. Design of Soil Sampling Programs:
Statistical Considerations. In: Quality Control in Reme-
dial Site Investigation: Hazardous and Industrial Solid
Waste Testing, 5th volume, C.L. Perket (ed.), ASTM
STP 925, American Society for Testing and Materials,
Philadelphia, PA, pp. 43-56.
Flatman, G.T. and A.A. Yfantis. 1984. Geostatistical Strategy
for Soil Sampling: The Survey and the Census. Environ-
mental Monitoring and Assessment 4:335-350.
Folorunso, O.A. and D.E. Rolston. 1984. Spatial Variability
of Field Measured Denitrification Gas Fluxes. Soil Sci.
Soc. Am. J. 48:1214-1219.
Fresenius, W., K.E. Quentin, and W. Schneider (eds.). 1988.
Water Analysis: A Practical Guide to Physico-Chemical
and Microbiological Water Examination and Quality As-
surance. Springer-Verlag, New York.
Gajem, Y. M., A. W. Warrick, and D.E. Myers. 1981. Spatial
Dependence of Physical Properties of a Typic Tornfluvent
Soil. Soil Sci. Soc. Am. J. 46:709-715.
Gibb, J.P, R.M. Schiller, and R.A. Griffin. 1981. Procedures
for the Collection of Representative Water Quality Data
from Monitoring Wells. ISWS/IGS Cooperative Ground
Water Report 7. Illinois State Water Survey, Champaign,
IL.
Gibbons, R.D. 1987a. Statistical Prediction Intervals for the
Evaluation of Ground-Water Quality. Ground Water
25(4):455-465.
Gibbons, R.D. 1987b. Statistical Models for the Analysis of
Volatile Organic Compounds in Waste Disposal Facili-
ties. Ground Water 25:572-580.
Gibbons, R.D. 1990. A General Statistical Procedure for
Ground-Water Detection Monitoring at Waste Disposal
Facilities. Ground Water 28(2):235-243.
Gilbert, R.O. 1987. Statistical Methods for Environmental
Pollution Monitoring. Van Nostrand Reinhold, New York.
Gilbert, R.O. and J.C. Simpson. 1985. Kriging from Estimat-
ing Spatial Pattern of Contaminants: Potential and Prob-
lems. Environmental Monitoring and Assessment
5:113-135.
Gillham, R. W., M.J.L. Robin, J.F. Barker and J.A. Cherry.
1983. Groundwater Monitoring and Sample Bias. API
Publication 4367. American Petroleum Institute, Wash-
ington, DC.
Gilhom, R.J. and D.R. Helsel. 1986. Estimation of Distribu-
tional Parameters for Censored Trace Level Water Qual-
ity Data: 1. Estimation Techniques. Water Resources
Research 22:135-146.
Gilliom, R.J., R.M. Hirsch, and E.J. Gilroy. 1984. Effect of
Censoring Trace-Level Water-Quality Data on Trend-
Detection Capability. Environ. Sci. Technol. 18:530-536.
Goerlitz, D.F. and E. Brown. 1972. Methods for Analysis of
Organic Substances in Water. U.S. Geological Survey
TWI 5-A3. (updated by Wershaw et al, 1987).
Harris, J., J.C. Loftis, and R.H. Montgomery. 1987. Statistical
Methods for Characterizing Ground-Water Quality.
Ground Water 25(2): 185-193.
Helsel, D.R. and R.J. Gilliom. 1986. Estimation of Distribu-
tional Parameters for Censored Trace Level Water Qual-
ity Data: 2. Verification and Applications. Water
Resources Research 22:146-155.
Hem, J.D. 1985. Study and Interpretation of the Chemical
Characteristics of Natural Water. U.S. Geological Survey
Water-Supply Paper 2254.
Ho, J.S-Y. 1983. Effect of Sampling Variables on Recovery
of Volatile Organics in Water. J. Am. Water Works Ass.
12:583-586.
Hollander, M. and D.A. Wolfe. 1973. Nonparametric Statisti-
cal Methods. John Wiley & Sons, New York.
99
-------
Holm, T.R. and C.D. Curtis. 1984. A Comparison of Oxida-
tion-Reduction Potentials Calculated from the As(V)/
As(III) and Fe(III)/Fe(II) Couples with Measured Plati-
num Electrode Potentials in Ground Water. J. Contami-
nant Hydrology 5:67-81.
Hughes, J.P. and D.P. Lettenmaier. 1981. Data Requirements
for Kriging: Estimation and Network Design. Water Re-
sources Research 17:1641-1650.
International Union of Pure and Applied Chemistry (IUPAC).
1978. Nomenclature, Symbols, Units and their Usage in
Spectrochemical Analysis—II. Data Interpretation.
Spectrochimica Acts B 33B:242-245.
Isaaks, E.H. and R.M. Srivastava. 1989. Applied Geostatistics.
Oxford University Press, New York.
Jenne, E.A. (ed.). 1979. Chemical Modeling in Aqueous
Systems: Speciation, Sorption, Volubility, and Kinetics.
ACS Symp. Series 93. American Chemical Society, Wash-
ington, DC.
lessen, R.J. 1978. Statistical Survey Techniques. John Wiley
& Sons, New York.
Journal, A.G. 1984. New Ways of Assessing Spatial Distribu-
tion of Pollutants. In: Environmental Sampling for Haz-
ardous Wastes, G.E. Schweitzer and J.A. Santolucito
(eds.), ACS Symp. Ser. 267, American Chemical Society,
Washington, DC, pp. 109-118.
Journal, A.G. and C.J. Huijbregts. 1978. Mining Geostatistics.
Academic Press, New York.
Kachonoski, R.G., D.E. Rolston, and E. deJong. 1985. Spatial
and Spectral Relationships of Soil Properties and
Microtopography: I. Density and Thickness of A Hori-
zon. Soil Sci. Soc. Am. J. 49:804-812.
Keith, S.J., M.T. Frank, G. McCarty, and G. Massman. 1983.
Dealing with the Problem of Obtaining Accurate Ground-
Water Quality Analytical Results. In: Proc. 3rd Nat.
Symp. on Aquifer Restoration and Ground Water Moni-
toring, National Water Well Association, Dublin, OH,
pp. 272-283.
Kirchmer, C.J. 1983. Quality Control in Water Analyses.
Environ. Sci. Technol. 17(4):174A-181A.
Kite, G. 1989. Some Statistical Observations. Water Re-
sources Bulletin 25(3):483-490. [See also 1990 discus-
sion by Kirby et al, and reply by Kite in Water Resources
Bulletin 26(4)693-698].
Klute, A. (ed.). 1986. Methods of Soils Analysis, Part 1-
Physical and Mineralogical Methods, 2nd ed. ASA Mono-
graph 9. American Society of Agronomy, Madison, WI.
Kopp, J.F. and G.D. McKee. 1983. Methods for Chemical
Analysis of Water and Wastes. EPA 600/4-74-020 (NTIS
PB84-128677). [supersedes report with the same title
dated 1979].
Lawrence, F.W. and S.B. Upchurch. 1976. Identification of
Geochemical Patterns in Groundwater by Numerical
Analysis. In: Advances in Groundwater Hydrology,
AWRA Proc. Ser. No. 21. American Water Resources
Association, Bethesda, MD.
Lehmann, E.L. 1975. Nonparametrics: Statistical Methods
Based on Ranks. Holden Day, San Francisco, CA, and
McGraw-Hill, New York.
Lm, P.C.L. 1986. User-Friendly IBM PC Computer Programs
for Solving Sampling and Statistical Problems. EPA/600/
4-86/023 (NTIS PB86-203783).
Long, G.L. and J.D. Winefordner. 1983. Limit of Detection, a
Closer Look at the IUPAC Definition. Analytical Chem-
istry 55(7):712A-724A.
Longbottom, J.E. and J.J. Lichtenberg. 1982. Methods for
Organic Chemical Analysis of Municipal and Industrial
Wastewater. EPA/600/4-82-057 (NTIS PB83-201798).
Mann, J.C. 1987. Misuses of Linear Regression in Earth
Sciences. In: Use and Abuse of Statistical Methods in the
Earth Sciences, W.B. Size (ed.), Oxford University Press,
New York, pp.74-106.
Matheron, G. 1971. The Theory of Regionalized Variables
and Its Applications. Cahiers du Centre de Morphologic
Mathematique de Fontainebleau. No. 5.
McBean, E.A. and F.A. Rovers. 1984. Alternatives for Han-
dling Detection Limit Data in Impact Assessments. Ground
Water Monitoring Review 4(2):42-44.
McBean, E.A. and F.A. Rovers. 1990. Flexible Selection of
Statistical Discrimination Tests for Field-Monitored Data.
In: Ground Water and Vadose Zone Monitoring, ASTM
STP 1053, D.M. Nielsen and A.I. Johnson (eds.), Ameri-
can Society for Testing and Materials, Philadelphia, PA,
pp. 256-265.
McBean, E. A., M. Kompter, and F. Rovers. 1988. A Critical
Examination of Approximations Implicit in Cochran's
Procedure. Ground Water Monitoring Review 8(1):83-
87.
McBratney, A.B. and R. Webster. 1981. Spatial Dependence
and Classification of the Soil Along a Transect in North-
east Scotland. Guoderma 26:63-82.
McNichols, R.J. and C.B. Davis. 1988. Statistical Issues and
Problems in Ground Water Detection Monitoring at Haz-
ardous Waste Facilities. Ground Water Monitoring Re-
view 8(4):135-150.
100
-------
Melchior, D.C. and R.L. Bassett (eds.). 1990. Chemical Mod-
eling of Aqueous Systems II. ACS Symp. Series 416.
American Chemical Society, Washington, DC.
Montgomery, R.H., J.C. Loftis, and J. Harris. 1987. Statistical
Characteristics of Ground Water Quality Variability.
Ground Water 25(2): 176-184.
Mueller, W., D.L. Smith, and L.H. Keith. 1991. Compilation
of EPA's Sampling and Analysis Methods. Lewis Pub-
lishers, Chelsea, MI, 465 pp. [On diskette EPA's Sam-
pling and Analysis Methods Database, Vol. 1 (Industrial
Chemicals), Vol. 2 (Pesticides, Herbicides, Dioxins and
PCBS), and Vol. 3 (Elements and Water Quality Param-
eters)].
Noblett, J.G. and J.M. Burke. 1990. FGD Chemistry and
Analytical Methods Handbook, 1: Process Chemistry—
Sampling, Measurement, Laboratory, and Process Perfor-
mance Guidelines, Revision 1. EPRI CS-3612. Electric
Power Research Institute, Palo Alto, CA. [Originally
published in 1984, see Radian Corporation (1988) for
Volume 2].
Olea, R.A. 1974. Optimal Contour Mapping Using Universal
Kriging. J. Geophysical Research 79(5): 696-702.
Olea, R.A. 1975. Optimum Mapping Techniques Using Re-
gionalized Variable Theory. KGS Series on Spatial Analy-
sis No. 2. Kansas Geological Survey, Lawrence, KS.
Ott, L. 1984. An Introduction to Statistical Methods and Data
Analysis, 2nd ed. Duxbury Press, Boston, MA, 775 pp.
Page, A.L., R.H. Miller, D.R. Keeney (eds.). 1982. Methods
of Soils Analysis, Part 2-Chemical and Microbiological
Properties, 2nd ed. ASA Monograph 9. American Society
of Agronomy, Madison, WI.
Pennine, J.D. 1988. There's No Such Thing as a Representa-
tive Ground Water Sample. Ground Water Monitoring
Review 8(3):4-9.
Porter, P.S. 1986. A Description of Measurement Error Near
Limits of Detection. In: Monitoring to Detect Changes in
Water Quality Series, D. Lerner (ed.), Int. Ass. of Hydro-
logical Sciences Pub. No. 157.
Porter, P.S., R.C. Ward, and H.F. Bell. 1988. The Detection
Limit. Environ. Sci. Technol. 22:856-861.
Provost L.P. and R.S. Elder. 1985. Choosing Cost-Effective
QA/QC Programs for Chemical Analysis. EPA/600/4-85/
056 (NTIS PB85-241461).
Radian Corporation. 1988. FGD Chemistry and Analytical
Methods Handbook, Vol. 2: Chemical and Physical Test
Methods, Revision 1. EPRI CS-3612. Electric Power
Research Institute, Palo Alto, CA. [Originally published
in 1984, see Noblett and Burke (1990) for Volume 1].
Rainwater, F.H. and L.L. Thatcher. 1960. Methods for Collec-
tions and Analysis of Water Samples. U.S. Geological
Survey Water-Supply Paper 1454.
Rice, G., J. Brinkman, and D. Muller. 1988. Reliability of
Chemical Analyses of Water Samples-The Experience
of the UMTRA Project. Ground Water Monitoring Re-
view 8(3):71-75.
Russo, D. and E. Bresler. 198la. Effect of Field Variability in
Soil Hydraulic Properties on Solutions and Unsaturated
Water and Salt Flows. Soil Sci. Soc. Am. J. 45:675-681.
Russo, D. and E. Bresler. 1981b. Soil Hydraulic Properties as
Stochastic Processes I. An Analysis of Field Spatial
Variability. Soil Sci. Soc. Am J. 45:682-687.
Samper, F.J. and S.P. Neuman. 1985. Geostatistical Analysis
of Hydrochemical Data from the Madrid Basin, Spain
(Abstract). Eos (Trans. Am. Geophysical Union)
66(46): 905.
Schuller, R.M., J.P. Gibb, and R.A. Griffin. 1981. Recom-
mended Sampling Procedures for Monitoring Wells.
Ground Water Monitoring Review l(l):42-46.
Seigel, S. 1956. Nonparametric Statistics for the Behavioral
Sciences, McGraw-Hill, New York.
Sgambat, J.P. and J.R. Stedinger. 1981. Confidence in Ground-
Water Monitoring. Ground Water Monitoring Review
l(Sprmg):62-69.
Sinclair, A.J. 1986. Statistical Interpretation of Soil Geo-
chemical Data. In: Exploration Geochemistry, Design
and Interpretation of Soil Surveys. Reviews in Economic
Geology 3:97-115.
Skoog, D.A. 1985. Principles of Instrumental Analysis, 3rd
ed. Saunders College Publishing, Philadelphia, PA.
Skougstad, M.W. et al. (eds.). 1979. Methods for Determina-
tion of Inorganic Substances in Water and Fluvial Sedi-
ments. U.S. Geological Survey TWI 5-A1.
Sisson, J.B. and P.J. Wierenga. 1981. Spatial Variability of
Steady-State Infiltration Rates as a Stochastic Process.
Soil Sci. Soc. Am. J. 45:699-704.
Smith, K.A. (ed.). 1991. Soil Analysis: Modem Instrumental
Methods, 2nd ed. Marcell Dekker, New York.
Sophodeous, M. I.E. Paschetto, and R.A. Olea. 1982. Gmund-
Water Network Design for Northwest Kansas, Using the
Theory of Regionalized Variables. Ground Water 20:48-
58.
Starks, T.H. and G.T. Flatman. 1991. RCRA Ground-Water
Monitoring Decision Procedures Viewed as Quality Con-
trol Schemes. Environmental Monitoring and Assess-
ment 16:19-37.
101
-------
Summers, K.V., G.L. Rupp, G.F. Davis, and S.A. Gherini.
1985. Ground Water Data Analysis at Utility Waste
Disposal Sites. EPRI EA-4165. Electric Power Research
Institute, Palo Alto, CA.
Thompson, C.M., et al. 1989. Techniques to Develop Data for
Hydrogeochemical Models. EPRI EN-6637. Electric
Power Research Institute, Palo Alto, CA.
Trangmar, B.B., R.S. Yost and G. Uehara. 1985. Application
of Geostatistics to Spatial Studies of Soil Properties.
Advances in Agronomy 38:45-93.
Tukey, J.W. 1977. Exploratory Data Analysis. Addison-
Wesley, New York, 506 pp.
U.S. Environmental Protection Agency (EPA). 1988. Field
Screening Methods Catalog: User's Guide. EPA/540/2-
88/005. FSMC System Coordinator, OERR, Analytical
Operations Branch (WH-548-A), U.S. EPA, Washington,
DC 20460.
U.S. Environmental Protection Agency (EPA). 1989. Guid-
ance Document on Statistical Analysis of Ground-Water
Monitoring Data at RCRA Facilities-Interim Final Guid-
ance. Office of Solid Waste Management Division (NTIS
PB89-151047).
van Kuilenburg, J., J.J. DeGruijter, B.A. Marsma, and J.
Bourna. 1982. Accuracy of Spatial Interpretation Be-
tween Point Data on Soil Moisture Supply Capacity
Compared with Estimations from Mapping Units.
Geoderma 27:311-325.
Vauclm, M., S.R. Vieira, R. Bernard, and J.L. Hatfield. 1982.
Spatial Variability of Surface Temperature Along to
Transects on a Bare Soil. Water Resources Research
18:1677-1686.
Vauclin, M., S.R. Vieira, G. Vachaud, and D.R. Nielsen.
1983. The Use of Co-Kriging with Limited Field Soil
Observations. Soil Sci. Soc. Am. J. 47:175-184.
Velleman, P.P. and D.C. Hoaglin. 1981. Applications, Basics,
and Computing of Exploratory Data Analysis. Duxbury
Press, Boston, MA, 354 pp.
Vieira, S.R., D.R. Nielsen, and J.W. Biggar. 1981. Spatial
Variability of Field Measured Infiltration Rate. Soil Sci.
Soc. Am. J. 45:1040-1048.
Warrick, A.W., D.E. Myers, and D.R. Nielsen. 1986. Geo-
statistical Methods Applied to Soil Science. In: Methods
of Soil Analysis, Part I—Physical and Mineralogical
Methods, 2nd ed., A. Klute (ed.), ASA Monograph No. 9,
American Society of Agronomy, Madison, WI, pp. 53-
Webster, R. 1977. Quantitative and Numerical Methods in
Soil Classification and Survey. Oxford University Press,
New York.
Wershaw, R.L., M.J. Fishman, R.R. Bragge, and L.E. Lowe
(eds.). 1987. Methods for the Determination of Organic
Substances in Water and Fluvial Sediments. U.S. Gee-
logical Survey TWI 5-A3. (Revision of Goerlitz and
Brown, 1972).
Westerman, R.L. 1990. Soil Testing and Plant Analysis, 3rd
ed. Soil Science Society of America, Madison, WI, 812
PP.
Willard, H. H., L.L. Memtt, Jr., J.A. Dean, and F.A. Settle, Jr.
1988. Instrumental Methods of Analysis, 7th ed.
Wadsworth Publishing Co., Belmont, CA.
Wilson, E.G., B.J. Adams, and B.W. Karney. 1990. Bias in
Log-Transformed Frequency Distributions. J. Hydrology
118:19-37.
Wollum II, A.G. and O.K. Cassel. 1984. Spatial Variability of
Rhizobium japonicum in Two North Carolina Soils. Soil
Sci. Soc. Am. J. 48:1082-1086.
Wood, W.W. 1976. Guidelines for Collection and Field Analy-
sis of Ground-Water Samples for Selected Unstable Con-
stituents. U.S. Geological Survey TWI 1-D2
Yates, S.R. and M.V. Yates. 1990. Geostatistics for Waste
Management: A User's Manual for the GEOPACK (Ver-
sion 1.0) Geostatistical Software System. EPA/600/8-90/
004 (NTIS PB90-186420/AS).
Yevjevich, V. and N.B. Harmancioglu. 1989. Description of
Periodic Variation in Parameters of Hydrologic Time
Series. Water Resources Research 25(3):421-428.
Yost, R.S., G. Uehara, and R.L. Fox. 1982. Geostatistical
Analysis of Soil Chemical Properties of Large Land
Areas. I. Variograms. Soil Sci. Sco. Am. J. 46:1028-
1032.
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Chapter 8
Geochemical Variability of the Natural and Contaminated Subsurface Environment
J. Russell Boulding and Michael J. Barcelona
This chapter focuses on subsurface geochemical pro-
cesses and environmental parameters that may significantly
affect the accuracy of geochemical sampling to characterize
the natural and contaminated subsurface. Subsequent chapters
examine in more detail subsurface physiochemical and deg-
radation processes that affect the fate and transport of con-
taminants. Table 8-1 indexes references on topics covered in
this chapter.
8.1 Overview of Subsurface Geochemistry
A basic assumption in performing remediation is that
one cannot remediate what is not observed. Consequently,
complete geochemical characterization of the subsurface re-
quires an understanding of what to observe and how to go
about making the observations. Elements and compounds in
the subsurface may exist in one or more of three phases (solid,
liquid, or gas). Within a phase, a substance may exist as
several forms or species (e.g., ions, neutral molecules, and
complex molecules in water). The partitioning of natural
constituents and contaminants between solid, liquid, and gas
or their transformation to other chemical forms is dependent
on both the thermodynamics and kinetics of different types of
chemical processes. Thermodynamic prediction and reaction
kinetics may be strongly influenced by subsurface environ-
mental conditions. Information on indicators of ground-water
conditions, such as pH, Eh, temperature, and pressure, there-
fore, is essential for interpreting geochemical data.
and whether a reaction will tend to occur. Thermodynamic
calculations can predict whether a chemical reaction is likely
to occur under specified conditions but give no indication of
how fast the reaction will occur. Kinetics describe the rate of
chemical reactions. Some reactions, such as the reaction that
occurs when a strong acid is added to water, will occur almost
instantaneously; other reactions, such as the hydrolysis of
cyanides at low pH, may take tens of thousands of years.
In nonequilibrium systems, chemical processes act to
alter the chemical composition and/or phase of the system,
and the system may tend to approach equilibrium. Simple
systems, such as dilute mixtures of sodium chloride and
water, attain solution equilibrium quickly, whereas complex
systems may only tend towards equilibrium. For example,
geochemical modeling by Apps et al. (1988) suggests that
Gulf Coast brines are not in equilibrium after tens of thou-
sands of years with respect to magnesium and sulfate concen-
tration. Lindberg and Runnells (1984) have suggested that
ground water is rarely, if ever, in complete equilibrium with
respect to redox reactions.
Equilibrium implies that as long as no significant changes
in environmental factors or phases occur within the system,
the chemical composition of the system will be predictable.
An equilibrium state does not imply that chemical reactions
cease, rather that the rates of forward and reverse reactions
compensate one other.
8.1.1 Geochemical Processes
Major geochemical processes in the subsurface include
(1) acid-base equilibria (also called ionization); (2) sorption-
desorption; (3) precipitation-dissolution; (4) oxidation-reduc-
tion (redox reactions); and (5) hydrolysis (see Chapters 10,
12, and 13). Microorganisms frequently are the catalysts or
promoters of reactions in the subsurface. Volatilization is
another important process affecting contaminants that readily
move into the gas phase. Interactions between these various
processes are typically complex and must be understood in
terms of both thermodynamic and kinetic controls.
Thermodynamically, a chemical system is in equilibrium
when its free energy is minimized; thus, thermodynamic
principles define the stability of substances within the system
8.1.2 Environmental Parameters
The act of sampling the subsurface tends to alter its
chemical equilibrium and results in reactions that may remove
or release some of the chemical constituents being measured.
The potential geochemical effects of drilling methods, materi-
als used for well construction and sampling devices, and
sampling methods all must be considered when developing a
sampling protocol. The sensitivity of a chemical system to
disturbance depends on a number of physical and chemical
environmental parameters. Some of the most important of
these parameters are discussed below, along with examples of
how sampling may bias the results of laboratory analyses.
The major geochemical parameters that characterize the
subsurface include (1) water content, (2) hydrogen ion con-
centration (pH), (3) redox potential (Eh), (4) microbial popu-
103
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Table 8-1. Sources of Information on Natural and Contaminant Variability of Geochemical Parameters In the Subsurface
Topic References
Soil Chemical Parameters
General chemistry
Background levels
Redox chemistry
Contaminants
Soil gases
Bohn et al. (1985), Bolt and Bruggenwert (1978), Dragun (1988), Fairbridge and Finkl (1979), Sparks (1986,
1989), Sposito (1984, 1989)
Connor and Shacklette (1975), Ebens and Shacklette (1982), Shacklette et al. (1971a,b, 1973, 1974)
Brookins (1988), Ponnamperuma (1972), Ransom and Smeck (1986)
Loehr et al. (1986)
Barber et al. (1990), van Cleemput and El-Sebaay (1985)
Soil Physical Parameters
Variability Jury (1985). See also Table 7-6
Flow channels Bouma et al (1983), Miller (1975), Simpson and Cunningham (1982), White (1985)
Vadose Zone
General
Water movement
Arnold et al. (1982), Evans and Nicholson (1987), Rijtema and Wassink (1969), Yaron et al. (1984), Zimmie and
Riggs (1979)
Barnes (1989), Diment and Watson (1985), Hill and Pariange (1972), Raats (1973)
Water Chemistry
General
Background levels
Redox Chemistry
Biochemical Changes
Corrosion/scaling
Variability
Drew (1989), Eriksson (1985), Faust and Aly (1981), Garrels and Christ (1965), Hem (1985), Lloyd and
Heathcote (1985), Morel (1983), Pagendorf (1978), Stumm and Morgan (1981)
Durum and Haffty (1961), Durum et al. (1971), Ebens and Shacklette (1982), Ledin et al. (1989), Leenheer et al.
(1974), Thurman (1985), White et al. (1963)
Baas-Backing et al. (1960), Back and Barnes (1965), Barcelona et al. (1989a), Champ et al. (1979), Edmunds
(1973), Hem and Cropper (1959), Lindberg and Runnells (1984), Smith et al. (1991), Zehnder and Stumm
(1988), ZoBell (1946). (See also, Tables 8-9 and 8- 10.)
Bouwer and McCarty (1984), Ghiorse and Wilson (1988), Smith et al. (1991), Wood and Bassett (1973)
Barnes and Clarke (1969), Langelier (1936), Larson and Buswell (1942), Ryzner (1944), Singley et al. (1985), Stiff
and Davis (1952).
Back and Hanshaw (1988), Montgomery et al. (1987), Schmidt (1977), Seaber (1965), van Beek and van Puffelen
(1987). (See also Tables 7-9 and 7-10.)
lation, (5) salinity and dissolved constituents, (6) physical and
chemical character of solids, (7) temperature, and (8) pres-
sure. Eh, pH, and pressure are probably the most important
parameters affecting sampling of near-surface aquifers; these
factors strongly influence microbial population. Dissolved
constituents and the physical and chemical character of sub-
surface solids are highly site specific and influenced primarily
by geologic and soil-forming processes. Salinity, temperature,
and solution composition gain increasing importance as the
depth of sampling increases.
pH and Alkalinity. The pH and alkalinity are master
variables that help to describe solution composition and po-
tential for precipitation reactions. For example, pump-and-
treat operations using air stripping to remove volatile organic
compounds (VOCs) can increase pH by 0.5 to 1 pH unit
through removing carbon dioxide, with subsequent precipita-
tion of calcium carbonate and iron oxides. Table 7-2 identifies
changes in pH that may result from sampling methods and
materials. Table 8-2 identifies the effects of pH on a number
of subsurface geochemical processes,
Alkalinity indicates the buffer capacity or resistance to
change in pH, A solution with high buffer capacity has a large
resistance to change in pH, requiring the addition of a propor-
tionally large amount of acid or base to change the solution
pH condition in the water. Since carbonate buffering is com-
mon to most natural waters, the solution pH may be quite
sensitive to volatilization of C02during sampling operations.
Redox Potential. The oxidation-reduction potential, or
Eh, is an expression of the intensity of redox conditions in a
system. It is measured in volts or millivolts (mV) as the
potential difference between a working electrode and the
standard hydrogen electrode. Positive readings in natural wa-
ter generally indicate oxidizing conditions, and negative read-
ings indicate reducing conditions. Ponnamperuma (1972)
suggests that Eh values of +200 mV or lower indicate reduc-
ing conditions in near-surface soils and sediments. Surface
water bodies are generally around 400 to 600 mV because
they are often in equilibrium with oxygen in the atmosphere.
Principal oxidizing species in ground-water systems are oxy-
gen and perhaps some hydrogen peroxide (the intermediate
species in the reduction of oxygen to water). Other oxidizing
species in ground water include nitrate and manganese (IV)
and Fe(III). Under reducing conditions, Fe(III) species will
tend to be reduced to Fe(II), sulfate is reduced to sulfide, and
104
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Table 8-2. Effects of pH on Subsurface Geochemical Processes and Other Environmental Factors
Process/Factor pH Effect
Acid-base
Adsorption-desorption
Precipitation-dissolution
Complexation
Oxidation-reduction
biodegradation
Measures acid-base reactions. Strong acids (bases) will tend to change pH: weak acids (bases) will buffer
solutions to minimize pH changes.
Strongly influences adsorption, because hydrogen ions play an active role in both chemical and physical bonding
processes. Abbility of heavy metals is strongly influenced by pH. Adsorption rates of organics are also PH
dependant.
Strongly influences precipitation-dissolution reactions. Mixing of solutions with different pH often results in
precipitation reactions. Sea also reservoir matrix below.
Strongly influences positions of equilibria involving complex ions and metal chelate formation.
Redox systems generally become more reducing with increasing pH (ZoBell, 1946).
in combination with Eh, strongly influences the types of bacteria that will be present. High-to medium-pH, low-Eh
environments will generally restrict bacterial populations to sulfate reducers and heterotrophic anaerobes
(Baas-Becking et a/., 1960).
Eh increasing pH generally lowers Eh.
Salinity pH-induced dissolution increases salinity: pH-induced precipitation decreases salinity.
Reservoir matrix Acidic solutions tend to dissolve carbonates and clays; highly alkaline solutions tend to dissolve silica and clays.
Greater pH generally increases cation-exchange capacity of clays.
Temperature
pH-driven exothermic (heat-releasing) reactions will increase fluid temperature: pH-driven endothermic (heat-
consuming) reactions will decrease fluid temperature.
Pressure Will not influence pressure unless pH-induced reactions result in a significant change in the volume of reaction
products.
Source: Adapted from U.S. EPA (1989)
carbon dioxide to methane. Oxidation/reduction processes are
discussed further in Section 12.1.3.
Most redox reactions in the subsurface are microbially
mediated. The measurement of the major by-products of these
reactions may be a better indicator of the strength of the
reducing environment than Eh measurements or calculated
equilibrium potentials. A sequence of redox reactions under
increasing reducing conditions may be (1) denitrification re-
actions which deplete nitrate and produce nitrogen gas, (2)
sulfate reduction which depletes sulfate and produces hydro-
gen sulfide, and (3) methanogenic reactions which deplete
carbon dioxide and produce methane. Microbially mediated
redox processes are discussed further in Section 12.2.3.
Redox potential measurements or calculated potentials
are only measures of intensity. Reduction capacity measures
the resistance to change in the redox potential, and is analo-
gous to buffer capacity for pH in water. Reduction capacity is
measured by how much oxidizing or reducing constituent
must be added to change redox conditions. Ground-water
systems tend to have some natural reduction capacity due to
the presence of organic carbon in aquifer solids. The introduc-
tion of organic contaminants, which serve as an energy source
for microorganisms to ground water, increases the tendency to
shift towards more reducing conditions. In contrast bias can
easily be introduced into analytical results by the addition of
oxygen during the sampling process. Increases in dissolved
oxygen, resulting in decreased Fe(II) concentrations in samples
(see Table 7-2), and precipitation of iron oxides are common
biases introduced by the exposure of ground-water samples to
the atmosphere.
The concept of biologically mediated redox zones is
useful for evaluating the biodegradation of organic contami-
nants in ground water. Table 8-3 shows how the degradation
of various organic micropollutants might occur with increas-
ing distance from a point of injection.
When organic contaminants are present in relatively low
concentrations, as with artificial recharge of treated sewage
effluent, oxygen is present near the zone of injection and
compounds susceptible to aerobic biodegradation will decom-
pose. As the redox potential declines at a greater distance
from the point of injection, denitrifying conditions develop,
and compounds such as carbon tetrachloride, which are not
105
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susceptible to aerobic degradation, may be degraded. If redox
potential declines further and conditions favorable for sulfate-
reducing bacteria exist, cresols and chlorophenols may be
degraded. Finally, where methanogenic bacteria predominate,
halogenated aliphatics that may have passed through the
denitrification zone may be degraded.
Implicit in this redox zone model is that compounds that
pass through the zone in which they are susceptible to biodeg-
radation will persist in ground water unless immobilized or
altered by inorganic chemical processes. In heavily contami-
nated ground water, this sequence may be reversed, with the
greatest reducing conditions closest to the point of contaminat-
ion grading to mildly oxygenated conditions (in shallow
aquifers, at least) at the outside edge of the contaminant
plume.
Salinity and Dissolved Constituents. Total dissolved sol-
ids (IDS) content can be qualitatively estimated in the field
by measuring specific conductance. The major dissolved con-
stituents in ground water may be near equilibrium with condi-
tions at their location, although subject to seasonal fluctuations
(see Section 8.4). During well development, if purging or
sampling-process ground water is mixed with water of differ-
ing salinity or chemical composition, the result may be pre-
cipitation-dissolution and redox reactions that significantly
change the inorganic chemistry of a sample. Geochemical
sampling of water wells that tap multiple aquifers is especially
problematic because of these effects. The more saline the
water, or the more different in chemical composition the two
waters, the greater the bias that can be introduced to geo-
chemical samples.
Soil/Aquifer Matrix. The mineralogy and particle size
distribution of the unsaturated and saturated zones strongly
influence geochemistry of subsurface waters. As particle size
decreases, the surface area increases, providing more opportu-
nities for chemical reactions between solids and water. A
particularly important chemical parameter of solids is the
cation exchange capacity (CEC). CEC is a function of miner-
alogy, particle size, and previous geochemical history. It may
be a good measure of the potential attenuation of pollutants by
ion exchange or sorption reactions. The CEC of clays is
strongly dependent on crystalline structure, with the high
shrink-swell smectite group (80 to 150 meq/100 g) having the
highest CEC and the nonswelling clays such as kaolinite the
lowest (3 to 15 meq/100 g). Characterization of clay mineral-
ogy can provide considerable insight into subsurface geo-
chemistry.
Temperature and Pressure. Temperature and pressure
directly influence the rate of chemical reactions. As pressure
increases, the amount of dissolved gases in solution tend to
increase. Consequently, sampling methods that allow gases
and VOCs to degas to the atmosphere at the land surface may
tend to underestimate concentrations. The deeper the sam-
pling, the greater the potential for errors resulting from pres-
sure changes.
Microbial Activity. Virtually all ground waters contain
diverse populations of microorganisms. The main limitation
to microbial growth in the subsurface is low levels of nutrient
and dissolved organic carbon. Microorganisms exist that are
capable of adapting to transform many types of organic con-
taminants. Unfortunately, most organic contaminants are more
readily degraded under aerobic conditions, and any contami-
nant loading that adds more than traces of contaminants will
rapidly deplete the available natural oxygen supply. As shown
in Table 8-3, halogenated aliphatic hydrocarbons and bromi-
nated methanes may be degraded under anaerobic conditions.
Phenols, alkyl phenols, and chlorophenols also may be de-
graded under these conditions (Wilson and McNabb, 1983).
Tetra- and trichloroethylene are readily degraded under
anaerobic conditions to intermediate daughter products, in-
cluding 1,2-dichloroethenes and 1,1 -dichloroethene, until vi-
nyl chloride is formed. Unfortumtely, vinyl chloride is resistant
to anaerobic degradation, although it readily degrades under
aerobic conditions. Other anaerobic degradation sequences
that end in relatively resistant compounds include carbon
tetrachloride to chloroform to methylene chloride and 1,1,1-
trichloroethane to 1,1-dichloroethane to chloroethane (Wood
et al, 1985).
Whether a specific contaminant will be degraded depends
on geochemical conditions and on the presence of microor-
ganisms that are capable of adaptation. Redox potential and
water chemistry can provide considerable insight into subsur-
face microbial activity even when samples are not taken for
microorganisms. Nitrogen, ammonia, hydrogen sulfide, and
methane in ground water are all indicators of microbial activ-
ity. Carbon dioxide also may indicate microbial activity;
however, its presence is more difficult to interpret because
carbon dioxide also may come from inorganic sources such as
calcium carbonate and dolomite. Section 13.2 discusses mi-
crobiological transformations in the subsurface in more detail.
8.1.3 The Vadose and Saturated Zones
The vadose and saturated zones have distinct geochemi-
cal differences that must be considered when sampling to
evaluate contamination. The vadose zone is a dynamic envi-
ronment with gases moving across the surface, the presence of
abundant organic matter, and solutes moving in and out of the
saturated zones. Gas transfers of interest include oxygen
going in, carbon dioxide moving out, and gases like nitrous
oxide or nitrogen being generated by bacteria. Organic matter
accumulation, weathering of minerals in the soil profile to
form clays, and the presence of air create a chemically reac-
tive environment.
The vadose zone also is characterized by considerable
heterogeneity in hydraulic conductivity. Macropores such as
old root channels, animal burrows, and channels between soil
structural units allow much more rapid movement of water
and associated contaminants than the aggregated soil particles
(see references in Table 8-1). These variations make represen-
tative sampling of soluble contaminants in the vadose zone
extremely difficult.
106
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Table 8-3. Redox Zones for Biodegradation of Organic Micropollutants
Increasing Distance from Injection Point —>
Biological Conditions
Aerobic
heterotrophic
respiration
Denitrification
Sulfate
respiration
Methanogenesis
Organic Pollutants Transformed
Chlorinated
benzenes
Ethylbenzene
Styrene
Naphthalene
Carbon tetrachloride
Bromodichloromethane
Dibromochloromethane
Bromoform
Phenol
Cresols
Chiorophenols
C,and C2
Halogenatad
aliphatics
Source: Adapted from Bouwer and McCarty (1984)
Table 8-4. Dissolved Solids in Potable Water -a Tentative
Classification of Abundance
Major Constituents (1.0 to 1000 ppm)
Sodium
Calcium
Magnesium
Silica
Bicarbonate
Sulfate
Chloride
Secondary Constituents (0.01 to 10.0 ppm)
iron
Strontium
Potassium
Boron
Carbonate
Nitrate
Fluoride
Minor Constituents (0.0001 to 0.1 ppm)
Antimony*
Aluminum
Arsenic
Barium
Bromide
Cadmium*
Chromium*
cobalt
Copper
Germanium*
iodide
Lead
Lithium
Manganese
Molybdenum
Nickel
Phosphate
Rubidium*
Selenium
Titanium*
Uranium
Vanadium
Zinc
Trace Constituents (generally <0.001 ppm)
Beryllium
Bismuth
Cerium*
Cesium
Gallium
Gold
Indium
Lanthanum
Niobium*
Platinum
Radium
Ruthenium*
Scandium*
Silver
Thallium*
Thorium*
Tin
Tungsten*
Ytterbium
Yttrium*
Zirconium*
* Element which occupies an uncertain position in the list.
Source: Adapted from Davis and DeWiest (1966)
8.2 Background Levels and Behavior of
Chemical Constituents
Interpretation of subsurface geochemical data requires
some knowledge of background levels as a baseline for evalu-
ating possible contamination and the chemical behavior of
individual constituents. Tables 8-4 and 8-5 show two classifi-
cation schemes for the abundance of dissolved species in
ground water. The first for potable water, includes only
dissolved solids and has four classes: major (1.0 to 1,000
ppm), secondary (0.1 to 10 ppm), minor (0.0001 to 0.1 ppm),
and trace (generally less than 0.001 ppm). The second scheme
is for highly mineralized water (> 1,000 mg/L), and includes
gases and organic acids. The classification of the organic
acids is based on data from the petroleum-bearing Frio forma-
tion in Texas (Kreitler et al., 1988). Organic acids for
nonpetroleum-bearing reeks would typically be in the minor
category.
Table 8-1 lists some sources of information on back-
ground levels of trace constituents in soils and ground water.
The U.S. Geological Survey is a good source of background
information on elemental composition of soils (Connor and
Shacklette, 1975; Ebens and Shacklette, 1982 and Shacklette
et al. 197 la,b, 1973, 1974) and water (Durum and Haffty,
1961; Durum et al. 1971; Ebens and Shacklette, 1982; White
et al., 1963). Thurman (1985), using data primarily from
Leenheer et al. (1974), reported the following median concen-
trations of organic carbon in various types of aquifers: sand
and gravel and limestone and sandstone - 0.7 mg/L; igneous -
0.5 mg/L; oil shales - 3.0 mg/L; organically rich recharge
waters - 10.0 mg/L, and petroleum associated wastes - 100
mg/L.
Table 8-1 also lists a number of general references on soil
and water chemistry and sources of information on more
specific geochemical topics such as redox chemistry, soil
gases, biochemical changes, and corrosion and sealing in
ground water. Tables 8-6 and 8-7 describe sources of informa-
tion on the chemical behavior of inorganic and organic natural
constituents and contaminants in the subsurface, respectively.
107
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Table 8-5. Classification of Dissolved Species in Deep- Water Injection Zones
Abundance Cations Anions Gases Organic Acids
Major
Sodium
Calcium
Magnesium
Chloride
Bicarbonate
Sulfate
Carbon Dioxide Acetate
Propionate
Intermediate
Minor
Silica
Barium
Potassium
Strontium
Boron
Iron"
Aluminun'
Manganes'
Arsenic
Beryllium
Cadmium
Chromium
cobalt
Copper
Lead
Lithium
Molybdenum
Nickel
Selenium
Zinc
Nitrate
Nitrite
Orthophosphate
Bromide
Iodide
Fluoride
Nitrogen
Hydrogen Sulfide
Methane
Butyrate
"Abundance classification criteria (mg/L): Major: 103-105Intermediate: 10'-103; Minor: <10'.
"Of possible special significance in assessing reactivity with injected wastes.
Source: U.S. EPA (1989)
8.3 Spatial Variability
Spatial variability of the subsurface is a result of scale
effects and physical and chemical gradients, which generally
exist both horizontally and vertically. Table 8-8 summarizes
typical ranges of subsurface environmental parameters that
may be found at a site. In general, contaminated sites have a
greater range of geochemical variation for all parameters than
do undisturbed sites. Spatial gradients for individual param-
eters are discussed below.
8.3.1 Scale
Soil and ground-water geochemistry vary regionally pri-
marily as a function of changes in climate and geology. An
important factor affecting ground-water chemistry is distance
from the recharge zone. In recharge zones, ground water tends
to be less mineralized than in areas of discharge. Regional-
scale changes in ground water are characterized by
hydrochemical facies (Seaber 1965); dominant chemical con-
stituents change with a shift in facies. Regional-scale patterns
in ground-water chemistry (e.g., Back and Hanshaw, 1971, on
carbonate equilibria) may not apply on the site scale. This is
particularly true with respect to oxygen-sensitive species,
because of disturbed land surface and substantial variability of
local recharge in surficial aquifers at the site level, which
influences oxygen concentration.
The maximum transport distance for contaminants de-
pends on the source and the medium of transport. Soil con-
tamination from atmospheric sources of heavy metals (lead,
zinc, cadmium) from smelters can extend from hundreds of
meters to kilometers. Contamination from underground stor-
age tanks (hydrocarbons and nonaqueous phase liquids
[NAPLs]) can have a radius of influence of about 50 to 2,000
m. NAPLs can migrate vertically 50 to 100 m.
8.3.2 Physical Gradients
Temperature Gradients. Temperature gradients affect
mixing, reaction paths and rate, and volubility. Vertical tem-
perature gradients can vary greatly, being very steep in geo-
thermal areas, but a good rule of thumb is that temperature
increases IT for every 50 to 60 feet of depth. Ground water
downgradient from a landfill may exhibit temperatures 8 to
12°F higher than water upgradient from a landfill.
Pressure Gradients. Vertical pressure gradients are on
the order of an atmosphere every 30 ft. Sampling mechanisms
used effectively at or near the land surface may not be valid
when used at 2 to 5 atmospheres (pressure at depths in excess
of 60 ft). Volatiles are in greater danger of being lost during
sampling when brought to the surface where the pressure is
lower.
108
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Table 8-6. Sources of Information on Chemical Behavior of Natural Inorganic Constituents and Contaminants in the Subsurface
Reference Description
Aubert and Pinta
(1977)
Bar-Yosef et al.
(1989)
Cattahan et al. (1979)
Copenhaver and
Wilkinson (1979)
Fdrstnerand Wittmann
(1979)
Fuller (1977)
Jacobs (1989)
Kabata-Pendias and
Pendias (1984)
Kramer and Duinker
(1984)
Lisk(1972)
McBride (1989)
Moore and Ramamoorthy
(1984a)
National Research
Council Canada (1982)
Nelson etal. (1983)
Purves (1977)
Rai and Zachara (1984)
Rai and Zachara (1988)
Singer (1973)
Thornton (1983)
Text on trace elements in soils. Includes chapters on
Bo, Cr, Co, Cu, I, Pb, Mn, Mo, Ni, Se, Ti, V, and Zn, and a chapter on 10 other minor elements (Li, Rb, Cs, Ba,
Sr, Bi, Ga, Ge, Ag, and Sn).
Collection of papers on behavior of inorganic contaminants in the vadose zone.
Data on environmental water-related fate of 129 organic and inorganic priority pollutants.
Bibliography with abstracts of articles from 1970 to 1974 on mobility of As, asbestos. Be, Cd, Cr, Cu,
cyanide, Pb, Hg, Se, and Zn in soil.
Comprehensive text on behavior of metal contaminants in the aquatic environment.
Review containing over 200 references on the movement of metals in soil.
Edited volume with 11 chapters on selenium in agriculture and the environment.
Text on trace elements in soils and plants.
Contains 42 papers on the complexation behavior of trace metals in natural waters.
Review paper on occurrence and chemistry of trace elements in soils and toxicities for plants and animals.
Review paper on reactions controlling heavy metal solubility in soils.
Book on behavior of heavy metals (As, Cd, Cr, Cu, Hg, Ni, andZn) in natural waters.
Data summary sheets on 16 selected toxic elements. NRCC reports on individual elements include:
chromium (1976), arsenic (1978a), lead (1978b), mercury (1979a), cadmium (1979b), and nickel (1981).
17 contributed chapters on chemical mobility and reactivity of the soil system with sections on principles of
chemical mobility and reactivity, biological activity and chemical mobility, and environmental impacts of toxic
chemical transport.
Text on trace element contamination of the environment focused primarily on soil.
Provides data on chemical attenuation rates, coefficients, and constants for 21 elements related to leachate
migration: Al, Sb, As, Ba, Be, B, Cd, Cr, Cu, F, Fe, Pb, Mn, Hg, Mo, Ni, Se, A/a, S, V, andZn.
Report containing laboratory data and equilibrium constants for key reactions needed to predict the geochemical
behavior of chromium in soil and ground water.
Contains 13 contributed chapters on trace metals and metal-organic interactions in natural waters.
Contains 16 contributed chapters on principles of environmental geochemistry with emphasis on heavy metals.
Velocity Gradients. Velocity gradients are a function of
pressure differences and hydraulic conductivity. Ground wa-
ter may flow at a rate of 10 to 100 m a day in the vicinity of
pumping wells. Increased velocity resulting from pumping
may have pronounced geochemical effects on ground-water
quality. Evidence of chemical zonation tends to be more
pronounced when water movement is rapid in relation to the
rate of chemical reactions (Baedecker and Back, 1979).
8.3.3 Chemical Gradients
A factor of 10 gradient in chemical concentrations over
vertical distances of less than 10 m is possible. Smith et al.
(1991) observed a 27-fold increase in bacterial abundance in a
9-m interval where an aquifer contained nitrate and organic
contaminants, Figure 8-1 shows changes in Eh, pH, oxygen,
and hydrogen sulfide in an aquifer from its point of outcrop-
ping where recharge occurs to about 24 km downdip. Dis-
solved oxygen dropped to zero about 11 km from the outcrop.
At the point that oxygen disappeared, Eh dropped signifi-
cantly from 400 mV to about 100 mV and continued to
decline slowly to around 0 mV at 24 km. Ground-water pH
showed a general upward trend. Once reducing conditions
prevailed in the aquifer, sulfate reduction, as evidenced by
hydrogen sulfide, was observed in 4 of the 10 sampling
points.
At the site level, redox potential can vary by a factor of 5
or 10 from the surface of a sand and gravel aquifer to a depth
of 100 feet in the same aquifer. Figure 8-2 shows vertical
changes in Eh, oxygen, and Fe(II). As in Figure 8-1, when
oxygen drops to zero at around 30 m, Eh drops and the
concentration of reduced Fe(II) increases dramatically. Tables
8-9 and 8-10 summarize examples of horizontal and vertical
rcdox gradients at the site and regional scales, respectively.
An uncontaminated aquifer may have a gradient in redox
potential of 30 or 50 mV/m vertically. In contaminated situa-
tions, redox potential may show a gradient of 150 mV/m
horizontally.
109
-------
Table 8-7. Sources of Information on Chemical Behavior of Natural Organic Constituents and Contaminants in the Subsurface
Reference
Description
Bitton and Gerba (1984)
Callahanetal.(1979)
Cheng (1990)
Faust and Hunter (1971)
Gerstl et al. (1989)
Gherinietal.(1988,
Ghiorse and Wilson
(1988)
Gibson (1984)
Goring and Hamaker
(1972)
Huang and Schnitzer
(1986)
Howard etal. (1991)
Kobayashi and Rittmann
(1982)
Lymanetal. (1982)
Mabeyetal. (1982)
Makietal. (1980)
Montgomery and Welkom
(1989), Montgomery
(1991)
Moore and Ramamoorthy
(1984b)
Morriletal. (1982)
Overcash (1981)
Sabljib (1987)
Sawhney and Brown
(1989)
Tabaketal. (1981)
Zehnder(1988)
Contains 14 papers focusing on the subsurface behavior of microorganisms as pollutants.
Data on environmental water-related fate of 129 organic and inorganic priority pollutants.
Collection of papers on the fate and transport of pesticides in soils.
Contains 24 papers on tf»e origin, occurrence, and behavior of organic compounds in aquatic environments.
Collection of papers on the fate of toxic organic chemicals in soil and ground water.
Compilation of data relevant to predicting the release, transport, transformation and fate of
more than 50 organic compounds.
Review of literature on biodegradation of organic contaminants in ground water.
Contains 15 papers on microbial degradation of organic compounds.
Two volumes containing 13 chapters on the behavior of organic chemicals in the soil environment.
Contains 15 contributed chapters on interactions of soil minerals with microbes and natural organic compounds.
Handbook of data on environmental degradation rates for more than 300 organic compounds.
Literature review summarizing about 90 examples of biodegradation of hazardous organic compounds.
Handbook on methods to estimate environmental behavior of organic compounds.
Aquatic fate process data for organic priority pollutants.
Contains 19 contributions to a workshop on biotransformation and fate of chemicals in the aquatic environment.
Volume 1 contains chemical data on 137 organic compounds commonly found in ground water and the
unsaturated zone. Volume 2 contains data on 267 additional compounds.
Book on behavior of organic chemicals (aliphatic hydrocarbons, mono- and polycyclic aromatic hydrocarbons,
chlorinated pesticides, petroleum hydrocarbons, phenols, PCBs and PCDD) in natural waters.
Text on sorption, degradation, and persistence of organic compounds in soils.
Contains 43 papers on decomposition of chlorinated organics, agricultural chemical, phenols, aromatic and
polynuclear aromatics, urea resins, and surfactants in soil.
Sorption coefficient data for 72 nonpolar and 159 polar and ionic organic compounds.
18 contributed chapters on reactions and movement of organic chemicals in soil.
Results of biodegradability studies for 114 organic priority pollutants.
Contains 14 papers on the biology of anaerobic microorganisms
Changes of a factor of four or five in pH, alkalinity, or
redox potential can mean magnitude changes in many chemi-
cal constituents. For example, in oxidizing conditions there is
virtually no dissolved iron in ground water. In anoxic ground
water reduced ferrous iron (Fe+2) can commonly approach 3 to
4 mg/L.
Samples from large-screen intervals in ground-water moni-
toring wells may give a misleading picture of subsurface
geochemistry as a result of mixing chemically different ground
waters. For example, Cowgill (1988) sampled a 10-m screen
by taking discrete grab samples from the top, middle, and
bottom of the screen interval and found that some metal
constituents differed by as much as a factor of 10.
8.4 Temporal Variability
Variations by a factor of two to five in the concentration
of the major ionic constituents (mg/L) in ground water can
occur for no apparent reason over the course of a hydrologic
year. Very little data are available for u/L level natural con-
stituents in ground water.
Shallow aquifers are particularly sensitive to changes in
pH and Eh in response to recharge events. Recharge at one
110
-------
Table 8-8. Ranges of Geochemically Significant Physical, Biological, and Chemical Values of Natural and Disturbed Near-Surface
Ground Water
Variable
Effects
Natural
Disturbed
Temperature
Pressure
Velocity
rnysicai vaiiauios
Mixing; reaction path and
rates: solubility
Gas solubility
Head differences/gradients
From pumping
Mixing from rapid
infiltration
3'-20'C
(61 0-1 5 'C)
1-10 bar
<1-10m/day
<1-1000m/day
3 '-35 'C
(MO-25'C)
1-1 000 bar
<1-100m/day
<1-1000m/day
Glucose
(Specific
activity)
Biological Variables
Biomass
Activity
Catalytic or transformation
potential
Turnover rates
Metabolic status
10'-10'cells/g
0. 1 \ng/Lhr
0.03-0.06x10'
104-10'cells/g
fig glucose/hrcell
Chemical Variables
pH
Conductance
Eh (mV)
Dissolved
Oxygen
(mg/L)
Alkalinity
(mg/L
CaCCy
See Table 7-2
Indicator of salinity
Redox status
Redox status
Buffer capacity
5.5 to 9.5
100 to 5000+
+600 to -100
<0.3 to 10
100 to 1000
3 to 12
100 to 10000+
+600 to -250
<0.3to>10
<100to>1000
time of the year may result in a set of chemical reactions
affecting chemical composition, whereas 6 months later an
entirely different set of reactions may occur. Thus, "represen-
tative" concentrations of background constituents may vary
seasonally.
Tables 8-11 and 8-12 summarize data on short-term
(minutes to days) and long-term (seasons to decades) varia-
tions of ionic constituents and several contaminants, respec-
tively. In general, both short-and long-term temporal variations
are less than an order of magnitude, with nitrate sometimes
showing a greater than order of magnitude variation (13X)
and Fe2+showing up to two orders of magnitude variation
(110X). Short-term variations generally result from individual
ground-water recharge events or well pumping and purging.
Seasonal variability generally results from variations in pre-
cipitation or irrigation, and multiyear trends typically result
from human activities such as salt-water intrusion from pump-
ing, irrigation, and fertilizer applications and nonagricultural
contamination.
Table 8-13 shows subjective estimates of strength of
seasonality or trend in 28 chemical constituents at three
different sites. The Sand Ridge site, which is far removed
from any sources of contamination, shows strong seasonal
trends in temperature and weak seasonal trends in alkalinity,
calcium, and magnesium concentrations. At the Beardstown
site, monitoring wells are located up- and down-gradient from
an anaerobic treatment lagoon for hog processing waste. The
contaminated downgradient wells at the Beardstown site ex-
hibited seasonality or trends for 16 constituents. The upgradient
ground water showed seasonality or trends for 12 constitu-
ents, an intermediate value between the pristine and contami-
nated ground water. For further information on references that
list methods for analyzing time-series water quality data for
seasonality and trend, see Table 7-7.
Ill
-------
• EH A O2
© pH * HS
0
•3
400
300 H
200
+ 100
0.
-100
EH
pH \ rt © © --"' ®
Pi1 -g-p © •** o
©
•
8.4
8.2
8.0
S •"
7.4
7.2
7.0
6-
4 -
2-
0-
Sample No. 1
i **"*•-•—•
A
/ \ A
X / \ / \
\ / V
•>•-A Ak^— A A-^A—A A A A^— A-
23 78 10 12 13 14 16 17 18 19
0.2
HS
0.1
10
20
Distance from Outcrop (km)
Start of
EH Sulfate
\Barrier Reduction Miyinnwith
' Mixing with
Connate Water
and Ion
Exchange
Piezometric Surface
of aquifer —j
345 • 78 • K> II IZ 15 14 IB 16 17 X 18
" Sea Level -J
Figure 8-1. Horizontal gradients in uncontaminated oxidation-reduction conditions, Lincolnshire limestone (from Champ et al.,
1979, after Edmunds, 1973).
112
-------
100 200 300 400 500 (mv) • Eh
2 4 6 8 10 (mg-L1) • Probe • Winkler, O,
10
20
O*.
30
0.1 0.2 0.3 0.4 0.5 (mg-L'1) A Fe*
Figure 8-2. Vertical gradients in uncontaminated oxidation-reduction conditions, Sand Ridge State Forest, Illinois (from
Barcelona et al., 1989a).
113
-------
Table 8-9. Spatial Gradients in Subsurface Oxidation-Reduction Conditions, Site Scale
Redox Gradient
Type of Environment
Unconfined sand
Unconfined sand
Unconfined
sand/gravel
Unconfined
sand/gravel
Unconfined sand
Confined sand/gravel
Confined sand/gravel
Unconfined sand
Unconfined sand
Unconiined
sand/gravel
Unconfined
sand/gravel
Unconfined sand
*Oy mg L< '
M '
-0.04
+0.1
-0.01
+0.5
-0.34"
-0.7
-0.2 to 0.77'
Eh, mV/m
Contaminant?
Horizontal (along general ground-water flow path)
+1 landfill leachate
-2 high organic carbon recharge
landfill leachate
-3 inorganic fertilizer plume
-1.5* anaerobic treatment leachate
-2.5 high organic carbon recharge water
artificial recharge
Vertical (increasing depth)
-10 to -15 background
-2 to -40 " landfill leachate
-30 high organic carbon recharge water
-2 to -30 c background
-8 to -27' anaerobic treatment leachate
Reference
Nicholson at a!. (1983)
Jackson and Patterson (1982)
Baedecker and Back (1979)
Barcelona and Naymik (1984)
this study (Baardsiown)
Jackson and Patterson (1982)
Van Beek and Van Puffelen
(1987)
Jackson et al. (1985)
Jackson et al. (1985)
Jackson and Patterson (1982)
this study (Sand Ridge)
this sutdy (Beardstown)
'Eighteen month average between wells 8 and 10.
bValues available from two separate sampling periods.
'Thirty month average range between wells 1 and 3 and 3 and 4, respectively.
Source: Barcelona et al., 1989a
Table 8-10. Spatial Gradients in Subsurface Oxidation-Reduction Conditions, Large Scale
Redox Gradient
Type of Gradient
Horizontal (along general
ground-water flow path)
Type of Environment
confined sandy clay/gravel (Patuxent)
confined sand/clay, lignite (Raritan-
Magothy)
confined carbonate chalk (Berkshire)
confined limestone (Lincolnshire)
confined sandstone/siltstone
(Foxhills-Basal Hell Creek)
Unconfined sand/gravel (Tucson
Basin)
'O., mg L-'
Km'
-0.30
-0.34
none
+1
AEh,mV/km
-34
-57
-30
-180
-0.4 to +5
+23
Reference
Back and Barnes (1965)
Back and Barnes (1965)
Edmunds etal. (1984)
Edmunds etal. (1984)
Thorstenson et al. (1979)
Rose and Long (1988)
Source: Barcelona et al., 1989a
114
-------
Table 8-11. Observations of Temporal Variations in Ground-Water Quality: Short-Term Variations
Nature of variability
Agricultural
Sources
Nonagricuiturai
or mixed sources
Constituents
(Concentration
variation)
Se(±2mg'L')
NO, - (1-3X)
SO4 = (3-7X)
NO> (1-4X)
NO, -(1-1 OX)
S04 = (1-1.5X)
NO3 - (0.5-2X)
Atrazine (1-5X)
H,S (1-5X)
SO4 = (1-1.2X)
NH3 (1-3X)
N03-(1-13X)
SO4 = (1-2X)
Fe (1-3X)
Mn(1-1.SX)
PCE, TCE, 1,2-t-DCE
(1-10X)
TCE(2-10X)
F&'(1-110X)
S=(1-15X)
Volatile halocarbons
(1-8X)
Period
Monthly
Minutes
Minutes
Monthly
Hours to
weeks
Minutes to
hours
Minutes to
hours
Minutes
Minutes
Monthly to
weekly
Minutes
Probable Cause
Irrigation/return/indeterminate
Pumpage/head changes and leaching
from unsaturated zone
Pumpage/vertical stratification
Irrigation/fertilizer applications/
leaching; locational differences
apparent
Surface runoff recharge
Pumping rate and well drilling
Pumping rate and purging
Purging
Purging rate and purging
Pumping rate and development
of cone of depression
Purging
Reference
Crist (1974)'
Schmidt (1977)'
Ecclesetal.(1977)"
Spalding and Exner (1980)
Libra etal. (1986)
Co!chineial.(1978)'
Humenicketal. (1980)'
Wilson and Rouse (1983)
Keely and Wolf (1983)'
McReynolds (1986)'
Barcelona and Helfrich (1986)
'Denotes variations observed in water supply production wells,
dichloroethylene
Source: Barcelona et a/., 1989b
PCE = perchloroethylene, TCE = trichloroethylene, 1,2-t-DCE =1,2 trans-
115
-------
Table 8-12. Observations of Temporal Variations in Ground- Water Quality: Long-Term Variations
Nature of variability
Agricultural
sources
Nonagricultural
or mixed sources
Constituents
(Concentration
variation)
CI-(+1.5X)
SO4 = (2-4X)
NO3-(3-6X)
SO4 = (3-7X)
NO3-(±48mg-L-'yr->)
NO,-(1-12X)
SO4 = (1- 1.5X)
NO3 - (1-5X)
NO3-(i-1.5X)
Pesticides (1-1.5X)
Conductance (2-3X)
SO4 = (1-3.5X)
Hardness (2-6X)
Conductance
(+2,000 \iS-cm-)
NO3-(±55mg>L-'yr->)
C1-(1-3X)
PCE±1-20X)
TCE(±1-3X)
Period
Decades
Seasonal
Seasonal
Seasonal
Seasonal
Years-seasonai
Seasonal
Decades
Seasonal
Seasonal
Seasonal
Seasonal
Probable Cause
Irrigation recharge
Irrigation/precipitation
Leaching/recharge
Irrigation/fertilizer applications
Recharge/fertilizer applications
infiltration/recharge
H2O level fluctuations
freezing/thawing recharge
Irrigation/upcoming of saline water
Sewage/fertilizer recharge and
applications
Oil field brine/recharge
Infiltrated surface water quality
variations
Pumping rate and patterns in
well field
Reference
Evenson (1965)'
Tenor/0 et al. (1969)'
Tryon (1976)
Spalding and Exner (1980)
Rajagopal and Talcott (1983)
Libra eiai. (1986)
Feulner and Schupp (1963)
Handy el at (1969)'
Perlmutter and Koch (1972)
Pettyjohn (1976, 1982)
Schwarzenbach et al. (1983)
McReynolds (1986)'
* Denotes variations observed in water supply production wells, PCE = perchloroethylene, TCE = trichloroethylene
Source: Barcelona et al., 1989b
116
-------
Table 8-13. Subjective Estimate of Strength of Seasonally or Trend Ground- Water Constituents In Uncontaininated (Sand Ridge
and Upgradient Beards town) and Contaminated (Downgradient Beardstown) Sites
Sand Ridge
(1-4)
pH
Cond
TempC +
TempW +
Eh
Probe O,
Wink O2
Alk
NH,
WO//
A/03A/0//
HS-
S04
SiO,
o-PO.
T-PO.
Cl-
Fe*
Ca
Mg
A/a
K
FeT
Mn,
TOX
VOC
NVOC
roc
Beardstown Beardstown Number of
(upgradient) (downgradient) violations
0
+ + 2
+ + 6
+ + 4
1
0
0
+ ° 1
3
1
0
0
0 0 0
0
7
1
+ 2
3
+ 1
2
° ° 3
° ° 3
0
+ 0
2
6
4
3
+ Indicates strongly seasonal.
"Indicates apparent trend or possible seasonality.
TOO = VOC + NVOC; Total Organic Carbon = Volatile Organic Carbon + Nonvolatile Organic Carbon.
Source: Barcelona et al., 1989b
8.5 References
Allen H., E.M. Perdue, and D. Brown (eds.). 1990. Metal
Speciation in Groundwater. Lewis Publishers, Chelsea,
MI.
Apps, J., L. Tsao, and O. Weres. 1988. The Chemistry of
Waste Fluid Disposal in Deep Injection Wells. In: Second
Berkeley Symposium on Topics in Petroleum Engineer-
ing, LBL-24337, Lawrence Berkeley Laboratory, Berke-
ley CA, pp. 79-82.
Arnold, E. M., G.W. Gee, and R.W. Nelson (eds.). 1982.
Proceedings of the Symposium on Unsaturated Flow and
Transport Modeling. NUREG/CP-0030. U.S. Nuclear
Regulatory Commission, Washington, DC.
Aubert, H and M. Pints. 1978. Trace Elements in Soils.
Elsevier, New York, 396 pp.
Baas-Becking, L.G.M., I.R. Kaplan, and D. Moore. 1960.
Limits of the Natural Environment in Terms of pH and
Oxidation-Reduction Potentials. J. Geology 68(3):243-
284.
Back, W. and I. Barnes. 1965. Relation of Electrochemical
Potentials and Iron Content to Ground Water Flow Pat-
terns. U.S. Geological Survey Professional Paper 498-C.
Back, W. and B. Hanshaw. 1971. Rates of Physical and
Chemical Processes in a Carbonate Aquifer. In: Non-
Equilibrium Concepts in Natural Water Chemistry, ACS
Adv. in Chemistry Series 106, American Chemical Soci-
ety, Washington, DC, pp. 77-93.
Baedecker, MJ. and W. Back. 1979. Modem Marine Sedi-
ments as a Natural Analog to the Chemically Stressed
Environment of a Landfill. J. Hydrology 43:393-414.
Barber, C., G.B. Davis, D. Briegel, and J.K. Ward. 1990.
Factors Controlling the Concentration of Methane and
Other Volatiles in Groundwater and Soil-Gas Around a
Waste Site. J. Contaminant Hydrology 5:155-169.
Barcelona, M.J. and J.A. Helfrich. 1986. Effects of Well
Construction Materials on Ground Water Samples.
Environ. Sci. Technol. 20(1 1): 1179-1184.
117
-------
Barcelona, M.J. and T.G. Naymik. 1985. Dynamics of a
Fertilizer Contaminant Plume in Ground Water. Environ.
Sci. Technol. 18(4):257-261.
Barcelona M.J., T.R. Helm, M.R. Schock, and O.K. George.
1989a. Spatial and Temporal Gradients in Aquifer Oxida-
tion-Reduction Conditions. Water Resources Research
25(5):991-1003.
Barcelona, M.J., D,P. Lettenmaier, and M.R. Schock. 1989b.
Network Design Factors for Assessing Temporal Vari-
ability in Ground-Water Quality. Environmental Moni-
toring and Assessment 12:149-179.
Barnes, C.J. 1989. Solute and Water Movement in Unsatur-
ated Soils. Water Resources Research 25(2):38-42.
Barnes, I. and F.E. Clarke. 1969. Chemical Properties of
Ground Water and Their Corrosion and Encrustation
Effects on Wells. U.S. Geological Survey Professional
Paper 498-D.
Bar-Yosef, B., N.J. Barrow, and J. Goldschmid (eds.), 1989.
Inorganic Chemicals in the Vadose Zone. Springer-Verlag,
New York.
Bitton, G. and C.P. Gerba (eds.). 1984. Groundwater Pollu-
tion Microbiology. Wiley-Interscience, New York.
Bohn, H.L., B.L. McNeal, and G.A. O'Connor. 1985. Soil
Chemistry, 2nd ed. Wiley-Interscience, New York.
Bolt, G.H. andM.G.M. Bruggenwert (eds.). 1978. Soil Chem-
istry. A. Basic Elements. Elsevier, New York, 282 pp.
Bouma, J., C. Belmans, L.W. Dekker, and W.J.M. Jeurissen.
1983. Assessing the Suitability of Soils with MacroPores
for Subsurface Liquid Waste Disposal. J. Environ. Qual.
12(3):305-311.
Bouwer, E.J. and P.L. McCarty. 1984. Modeling of Trace
Organics Biotransformation in the Subsurface. Ground
Water 22:433-440.
Brookins, D.G. 1988. Eh-pH Diagrams for Geochemistry.
Springer-Verlag, New York, 176 pp.
Callahan, M.A. et al. 1979. Water-Related Environmental
Fate of 129 Priority Pollutants (2 Volumes). EPA/440/4-
79-029a-b.
Champ, D.R., J. Gulens, and R.E. Jackson, 1979. Oxidation-
Reduction Sequences in Ground Water Flow Systems.
Can. J. Earth Sci. 16:12-23.
Cheng, H.H. (ed.). 1990. Pesticides in the Soil Environment:
Processes, Impacts and Modeling, Soil Science Society
of America, Madison, WI, 554 pp.
Colchin, M.P., LJ. Turk, and MJ. Humenick. 1978. Sam-
pling of Ground Water Baseline and Monitoring Data for
In Situ Processes. Water Resources Research Center Re-
port EHE 78-01, University of Texas, Austin, TX.
Connor, J.J. and H.T. Shacklette. 1975. Background Geo-
chemistry of Some Rocks, Soils, Plants, and Vegetables
in the Conterminous United States. U.S. Geological Sur-
vey Professional Paper 574-F.
Copenhaver, E.D. and B.K. Wilkinson. 1979. Movement of
Hazardous Substances in Soil: A Bibliography, Vol. I:
Selected Metals, Municipal. EPA/600/9-79-024a (NTIS
PB80-113103).
Cowgill, U. 1988. Sampling Waters: The Impact of Sample
Variability on Planning and Confidence Levels. In Prin-
ciples of Environmental Sampling, L.H. Keith (ed.), ACS
Professional Reference Book, American Chemical Soci-
ety, Washington, DC, Chapter 11.
Crist, M.A. 1974. Selenium in Waters in and Adjacent to
Kendrick Project, Natrona County, WY. U.S. Geological
Survey Water-Supply Paper 2023.
Davis, S.N. and RJ. DeWiest 1966. Hydrogeology. John
Wiley & Sons, New York.
Diment, G.A. and K.K. Watson. 1985. Stability Analysis of
Water Movement in Unsarurated Porous Materials 3.
Experimental Studies. Water Resources Research
21(7):979-984.
Dragun, J. 1988. The Soil Chemistry of Hazardous Materials.
Hazardous Materials Control Research Institute, Silver
Spring, MD, 447+ pp.
Drever, J.I. 1989. The Geochemistry of Natural Water, 2nd
ed. Prentice-Hall, Englewood Cliffs, NJ.
Durum, W .H. and J. Haffty. 1961. Occurrence of Minor
Elements in Water. U.S. Geological Survey Circular 445.
Durum, W.H., J.D. Hem, and S.G. Heidel. 1971. Reconnais-
sance of Selected Minor Elements in Surface Waters of
the United States, October 1970. U.S. Geological Survey
circular 643.
Ebens, R.J. and H.T. Shacklette. 1982. Geochemistry of Some
Rocks, Mines Spoils, Stream Sediments, Soils, Plants and
Waters in the Western Energy Region of the Contermi-
nous United States. U.S. Geological Survey Professional
Paper 1238.
Eccles, L. A., J.M. Klein, and W.F. Hardt. 1978. USGS Scien-
tists Bring California Water Supply into Compliance with
Federal Regulations. Water Well Journal 13(2):42-45.
Edmunds, W.M. 1973. Trace Element Variations Across an
Oxidation-Reduction Barrier in a Limestone Aquifer. In
Proc. Symp. on Hydrogeochemistry and Biogeochemis-
tg (Tokyo, 1970), E. Ingerson (ed.), Clarke Company,
Washington, DC, pp. 500-528.
Edmunds, W.M., D.L. Miles, and J.M. Cook. 1984. A Com-
parative Study of Sequential Redox Processes in Three
British Aquifers. In: Hydrochemical Balances of Fresh-
118
-------
water Systems, Int. Ass. of Hydrological Sciences Pub.
No. 150, pp. 55-70.
Eriksson, E. 1985. Principles and Applications of
Hydrochemistry. Chapman and Hall, New York.
Evans, D.D. and T.J. Nicholson (eds.). 1988. Flow and Trans-
port Through Unsaturated Fractured Rock. AGU Geo-
physical Monograph 42. American Geophysical Union,
Washington, DC.
Evenson, R.E. 1965. Suitability of Irrigation Water and
Changes in Ground-Water Quality in the Lompoc Subarea
of the Santa Ynez River Basin, Santa Barbara, CA. U. S.
Geological Survey Water-Supply Paper 1809-S.
Fairbridge, R.W. and C.W. Finkl, Jr. (eds.). 1979. The
Encylcopedia of Soil Science, Part 1: Physics, Chemistry,
Biology, Fertility, and Technology. Dowden, Hutchinson
& Ross, Stroudsburg, PA, 646 pp.
Faust, S.D. and O.M. Aly. 1981. Chemistry of Natural Wa-
ters. Ann Arbor Science Publishers, Ann Arbor, MI.
Faust, S.D. and J.V. Hunter (eds.). 1971. Organic Compounds
in Aquatic Environments. Marcel Dekker, New York.
Feulner, A.J. and R.G. Schupp. 1963. Seasoml Changes in the
Chemical Quality of Shallow Ground Water in North-
western Alaska. U.S. Geological Survey Professional
Paper 475-B, pp. B189-B191.
Forstner, U. and G.T.W. Wittmann. 1979. Metal Pollution in
the Aquatic Environment. Springer-Verlag, New York.
Fuller, W.H. 1978. Movement of Selected Metals, Asbestos
and Cyanide in Soils: Applications to Waste Disposal
Problems. EPA/600/2-77-020 (NTIS PB 266905).
Garrels, R.M. and C.L. Christ. 1965. Solutions, Minerals and
Equilibria. Freeman, Cooper & Co., San Francisco, CA.
Gerstl, Z., Y. Chen, U. Mmgelgrm, and B. Yaron (eds.). 1989.
Toxic Organic Chemicals in Porous Media. Springer-
Verlag, New York.
Ghermi, S. A., K.V. Summers, R.K. Munson, and W.B. Mills.
1988. Chemical Data for Predicting the Fate of Organic
Compounds in Water, Vol. 2: Database. EPRI EA-5818.
Electric Power Research Institute, Palo Alto, CA.
Ghermi, S.A., K.V. Summers, R.K. Munson, and W.B. Mills.
1989. Chemical Data for Predicting the Fate of Organic
Compounds in Water, Vol. 1: Technical Basis. EPRI EA-
5818. Electric Power Research Institute, Palo Alto, CA.
Ghiorse, W.C. and J.T. Wilson. 1988. Microbial Ecology of
the Terrestrial Subsurface. Adv. Appl. Microbiol. 33:107-
172.
Gibson, D.T. (ed.). 1984. Microbial Degradation of Organic
Compounds. Marcel Dekker, New York.
Goring, C.A.I, and J.W. Hamaker. 1972. Organic Chemicals
in the Soil Environment, 2 Volumes. Marcel Dekker,
New York.
Handy, A.H., R.W. Mower, and G.W. Sandberg. 1969. Changes
in the Chemical Quality of Ground Water in Three Areas
in the Great Basin, UT. U.S. Geological Survey Profes-
sional Paper 650-D, pp. D228-D234.
Hem, J.D. 1985. Study and Interpretation of the Chemical
Characteristics of Natural Water, 3rd ed. U.S. Geological
Survey Water-Supply Paper 2254.
Hem, J.D. and W.H. Croper. 1959. Survey of Ferrous-Ferric
Chemical Equilibria Redox Potentials. U.S. Geological
Survey Water-Supply Paper 1959-A.
Hill, D.E. and J.-Y. Parlange. 1972. Wetting Front Instability
in Layered Soils. Soil Sci. Sco. Am. Proc. 36(5):697-702.
Howard, P. H., W.F. Jarvis, W.M. Meylan, and E.M.
Mikalenko. 1991. Handbook of Environmental Degrada-
tion Rates. Lewis Publishers, Chelsea, MI, 700+ pp.
Huang, P.M. and M. Schnitzer. 1986. Interactions of Soil
Minerals with Natural Organics and Microbes. SSSA Sp.
Pub. No. 17. Soil Science Society of America, Madison.
WI, 606 pp.
Humemck, M.J., L.J. Turk, and M.P. Colchm. 1980. Method-
ology for Monitoring Ground Water at Uranium Solution
Mines. Ground Water 18(3): 262-273.
Jackson, RE. and R.J. Patterson. 1982. Inteq)retation of pH
and Eh Trends in a Fluvial-Sand Aquifer System. Water
Resources Research 18(4): 1255-1268.
Jackson, R.E., R.J. Patterson, B.W. Graham, J. Bahr, D.
Belanger, J. Lockwood, and M. Priddle. 1985. Contami-
nant Hydrogeology of Toxic Organic Chemicals at a
Disposal Site, Gloucester, Ontario. 1. Chemical Concepts
and Site Assessment. National Hydrologic Research In-
stitute Paper 23. Canadian Department of Environment,
Inland Water Branch, Ottawa, Ontario.
Jacobs, L.W. (ed.). 1989. Selenium in Agriculture and the
Environment. SSSA Sp. Pub. No. 23. Soil Science Soci-
ety of America, Madison, WI, 233 pp.
Jury, W.A. 1985. Spatial Variability of Soil Physical Param-
eters in Solute Migration: A Critical Literature Review.
EPRI EA-4228. Electric Power Research Institute, Palo
Alto, CA.
Kabata-Pendias, A. and H. Pendias. 1984. Trace Elements in
Soils and Plants. CRC Press, Boca Raton, FL, 336 pp.
Keely, J.F. and F. Wolf. 1983. Field Applications of Chemical
Time-Series Sampling. Ground Water Monitoring Re-
view 3(4):26-33.
119
-------
Kobayashi, H. and B.E. Rittmann. 1982. Microbial Removal
of Hazardous Organic Compounds. Environ. Sci. Technol.
16:170A-183A.
Kramer, C.J.M. and J.C. Duinker. 1984. Complexation of
Trace Metals in Natural Waters. Martinus Nijhoff/Dr W.
Junk Publishers, Boston.
Kreitler, C.W., M.S. Akhter, and A.C.A. Donnelly. 1988.
Hydrogeologic-Hydrochemical Characterization of Texas
Gulf Coast Formations Used for Deep-Well Injection of
Chemical Wastes. Bureau of Economic Geology, Univer-
sity of Texas-Austin, TX.
Langelier, W .F. 1936. The Analytical Control of Anti-Corro-
sion Water Treatment. J. Am. Water Works Ass. 28:1500-
1521.
Larson, I.E. and A.M. Buswell. 1942. Calcium Carbonate
Saturation Index and Alkalinity Interpretations. J. Am
Water Works Ass. 34:1667-1684.
Ledin, A., C. Pettersson, B. Allard, and M. Aastrup. 1989.
Background Concentration Ranges of Heavy Metals in
Swedish Groundwaters from Crystalline Rocks: A Re-
view. Water, Air, and Soil Pollution 47:419-426. In-
cludes: Cr. Cu, Zn, Cd, Pb.
Leenheer, J.A., R.L. Malcolm, P.W. McKinley, and L.A.
Eccles. 1974. Occurrence of Dissolved Organic Carbon
in Selected Groundwater Samples in the United States. J.
Res. U.S. Geological Survey 2:361-369.
Libra, R.D., G.R. Hallberg, B.E. Hoyer, and L.G. Johnson.
1986. Agricultural Impacts on Ground Water Quality:
The Big Spring Basin Study, Iowa. In: Proc. of Agricul-
tural Impacts on Ground Water (Omaha, NB), National
Water Well Association, Dublin, OH, pp. 252-273.
Lindberg, O.K. and D.D. Runnells. 1984. Ground Water
Redox Reactions: An Analysis of Equilibrium State Ap-
plied to Eh Measurements and Geochemical Modeling.
Science 225:925-928.
Lisk, D.J. 1972. Trace Metals in Soils, Plants and Animals.
Advances in Agronomy 24:267-325.
Lloyd, J. W, and J.A. Heathcote. 1985. Natural Inorganic
Hydrochemistry in Relation to Groundwater. Oxford Uni-
versity Press, New York.
Loehr, R.C., J.H. Martin, Jr., E.F. Neuhauser. 1986. Spatial
Variation of Characteristics in the Zone of Incorporation
at an Industrial Waste Land Treatment Site. In: Hazard-
ous and Industrial Solid Waste Testing: Fourth Sympo-
sium, ASTM STP 886, J.K. Petros, Jr., W.J. Lacy, and
R.A. Conway (eds.), American Society for Testing and
Matenals, Philadelphia, PA, pp. 285-298.
Lyman, W.J., W.F. Reehl, and D.H. Rosenblatt (eds.). 1982.
Handbook of Chemical Property Estimation Methods:
Environmental Behavior of Organic Compounds.
McGraw-Hill, New York.
Mabey, W.R., et al. 1982. Aquatic Fate Process Data for
Organic priority Pollutants. EPA/440/4-81-014 (NTIS
PB87-169090).
Maki, A.W., K.L. Dickson, and J. Cairns, Jr. (eds.). 1980.
Biotransformation and Fate of Chemicals in the Aquatic
Environment. American Society for Microbiology, Wash-
ington, DC.
McBride, M.A. 1989. Reactions Controlling Heavy Metal
Volubility in Soils. In Advances in Soil Science, B.A.
Stewart (ed.), Springer-Verlag, New York, Vol. 10.
McReynolds, L. 1986. Monitoring Organic Contaminants in
Los Angeles San Fernando Valley Ground Water Basin.
In: Proc. of the 15th Biennial Conference on Ground
Water, J.J. Devries (ed.), California Water Resources
Center, San Diego, CA, pp. 53-60.
Miller, E.E. 1975. Physics of Swelling and Cracking Soils. J.
Colloid and Interface Science 52(3):434-443.
Montgomery, J.H. 1991. Ground Water Chemical Desk Ref-
erence, Vol. 2. Lewis Publishers, Chelsea, ML
Montgomery, J.H. and L.M. Welkom. 1989. Ground Water
Chemicals Desk Reference. Lewis Publishers, Chelsea,
MI.
Montgomery, R.H., J.C. Loftis, and J. Harris. 1988. Statistical
Characteristics of Ground Water Quality Variability.
Ground Water 25(2): 176-184.
Moore, J.W. and S. Ramamoorthy. 1984a. Heavy Metal in
Natural Waters: Applied Monitoring and Impact Assess-
ment. Springer-Verlag, New York.
Moore, J.W. and S. Ramamoorthy. 1984b. Organic Chemicals
in Natural Waters: Applied Monitoring and Impact As-
sessment. Springer-Verlag, New York.
Morel, F.M.M. 1983. Principles of Aquatic Chemistry. Wiley
Interscience, New York, 435+ pp.
Morril, L.G., B. Mahalum, and S.H. Mohiuddin. 1982. Or-
ganic Compounds in Soils: Sorption, Degradation and
Persistence. Ann Arbor Science/The Butterworth Group,
Wobum, MA, 326 pp.
National Research Council Canada, 1976. Effects of Chro-
mium in the Canadian Environment, NRCC Report No.
15018. Ottawa, Ontario.
National Research Council Canada, 1978a. Effects of Arsenic
in the Canadian Environment. NRCC Report No. 15391.
Ottawa, Ontario.
National Research Council Canada, 1978b. Effects of Lead in
the Environment - 1978: Quantitative Aspects. NRCC
Report No. 16736. Ottawa, Ontario.
120
-------
National Research Council Canada, 1979a. Effects of Mer-
cury in the Canadian Environment. NRCC Report No.
16739. Ottawa, Ontario.
National Research Council Canada, 1979b. Effects of Cad-
mium in the Canadian Environment. NRCC Report No.
16743. Ottawa, Ontario.
National Research Council Canada, 1981. Effects of Nickel in
the Canadian Environment. NRCC Report No. 18568
(Reprint). Ottawa, Ontario.
National Research Council Canada. 1982. Data Sheets on
Selected Toxic Elements. NRCC Report No. 19252. Ot-
tawa, Ontario. Includes: Sb, Ba, Be, Bi, B, Cs, Ga, Ge, In,
Mo, Ag, Te, Tl, Sn (inorganic and organic), U, Zr.
Nelson, D.W., D.E. Elrick, and K.K. Tanji. 1983. Chemical
Mobility and Reactivity in Soil Systems. SSSA Sp. Pub.
No. 11. Soil Science Society of America, Madison, WI,
262 pp.
Nicholson, R. V., J.A. Cherry and E.J. Reardon. 1983. Migra-
tion of Contaminants in Ground Water at a Landfill, a
Case Study: 6. Hydrogeochemistry. J. Hydrology 63:131-
176.
Overcash, M.R. (ed.). 1981. Decomposition of Toxic and
Nontoxic Organic Compounds in Soil. Ann Arbor Sci-
ence/The Butterworth Group, Wobum, MA, 375 pp.
Pagendorf, G.K. 1978. Introduction to Natural Water Chemis-
try. Marcel Dekker, New York.
Perlmutter, N.M. and E. Koch. 1972. Preliminary Hydrogeo-
logic Appraisal of Nitrate in Ground Water and Streams,
Southern Nassau County, Long Island, NY. U.S. Geo-
logical Survey Professional Paper 800-B, pp. B225-B235.
Pettyjohn, W.A. 1976. Monitoring Cyclic Fluctuations on
Ground-Water Quality. Ground Water 14(6):472480.
Pettyjohn, W.A. 1982. Cause and Effect of Cyclic Changes in
Ground-Water Quality. Ground Water Monitoring Re-
view 2(l):43-49.
Ponnamperuma, F.N. 1972. The Chemistry of Submerged
Soils. Advances in Agronomy 24:29-98.
Purves, D. 1978. Trace-Element Contamination of the Envi-
ronment. Elsevier, New York.
Raats, P.A.C. 1973. Unstable Wetting Fronts in Uniform and
Nonuniform Soils. Soil Sci. Sco. Am. Proc. 37:681-685.
Rai, D. and J.M. Zachara. 1984. Chemical Attenuation Rates,
Coefficients and Constants in Leachate Migration. Vol.
1: A Critical Review. Vol. 2: Annotated Bibliography,
EPRI EA-3356. Electric Power Research Institute, Palo
Alto, CA.
Rai, D. and J.M. Zachara. 1988. Chromium Reactions in
Geologic Materials. EPRI EA-5741. Electnc Power Re-
search Institute, Palo Alto, CA.
Rajagopal, R. and R.L. Talcott. 1983. Patterns in Groundwa-
ter Quality Selected Observations in Iowa. Environmen-
tal Management l(5):465-474.
Ransom, M.D. andN.E. Smeck. 1986. Water Table Charac-
teristics and Water Chemistry of Seasonally Wet Soils of
Southwestern Ohio. Soil Sci. Sco. Am. J. 50:1281-1290.
Rijtema, P.E., and H. Wassink (eds.). 1969. Water in the
Unsaturated Zone, 2 Volumes. Studies and Reports in
Hydrology No. 2, UNESCO, Pans.
Ryzner, J.W. 1944. A New Index for Determining Amount of
Calcium Carbonate Scale Formed by Water. J. Am. Wa-
ter Works Ass. 36:472-486.
Rose, S. and A. Long. 1988. Dissolved Oxygen Systematic in
the Tucson Basin Aquifer. Water Resources Research
24(1):127-136.
Sablic, A. 1988. On the Prediction of Soil Sorption Coeffi-
cients of Organic Pollutants by Molecular Topology.
Environ. Sci. Technol, 21(4):358-366.
Sawhney, B.L. and K. Brown (eds.). 1989. Reactions and
Movement of Organic Chemicals in Soils. SSSA Special
Publ. No. 22. American Society of Agronomy, Madison,
WI.
Schmidt, K.D. 1978. Water Quality Variations for Pumping
Wells. Ground Water 15(2): 130-138.
Schwarzenbach, R.P., W. Giger, E. Hoehn, and J.K. Schneider.
1983. Behavior of Organic Compounds During Infiltra-
tion of River Water to Ground Water Field Studies.
Environ. Sci. Technol. 17(8):472-479.
Seaber, P.R. 1965. Variations in Chemical Character of Water
in the Englishtown Formation, New Jersey. U.S. Geo-
logical Survey Professional Paper 498-B.
Shacklette, H.T. et al. 1971a. Elemental Composition of Surfi-
cial Materials in the Conterminous United States. U.S.
Geological Survey Professional Paper 574-D. Includes:
Al, Ba, Be, Bo, Ca, Ce, Cr, Co, Cu, Ga, Fe, La, Pb, Mg,
Mo, Ne, Ni, Nb, P, K, Sc, Na, Sr, Ti, V, Y, Yb, Zn, Zr.
Shacklette, H.T. et al. 1971b. Mercury in the Environment—
Surficial Materials of the Conterminous United States.
U.S. Geological Survey Circular 644.
Shacklette, H.T. et al. 1973. Lithium in Surficial Materials of
the Conterminous United States and Partial Data on Cad-
mium. U.S. Geological Survey Circular 673.
Shacklette, H.T. et al. 1974. Selenium, Fluorine, and Arsenic
in Surficial Materials of the Conterminous United States.
U.S. Geological Survey Circular 692.
121
-------
Simpson, T.W. and RL. Cunningham. 1982. The Occurrence
of Flow Channels in Soils. J. Environ. Qual. 11(1):29-30.
Singer, P.C. 1973. Trace Metals and Metal Organic Interac-
tions in Natural Waters. Ann Arbor Science, Ann Arbor,
MI.
Singley, I.E., R.A. Pisigan, Jr., A. Ahmadi, P.O. Pisigian, and
T-Y. Lee. 1985. Corrosion and Calcium Carbonate Satu-
ration Index in Water Distribution Systems. EPA/600/2-
85/079 (NTISPB85-228112)
Smith, R.L., R.W. Harvey, and D.R. LeBlanc. 1991. Impor-
tance of Closely Spaced Vertical Sampling in Delineating
Chemical and Microbiological Gradients in Groundwater
Studies. J. Contaminant Hydrology 7:285-300.
Spalding, R.F. and M.E. Exner. 1980. Areal, Vertical and
Temporal Differences in Ground Water Chemistry: I.
Inorganic Constituents. J. Env. Quality 9(3):466-479.
Sparks, D.L. (ed.). 1986. Soil Physical Chemistry. CRC Press,
Boca Raton, FL, 320 pp.
Sparks, D.L. 1989. The Kinetics of Soil Chemical Processes.
Academic Press, New York.
Sposito, G. 1984. The Surface Chemistry of Soils. Oxford
University Press, New York.
Sposito, G. 1989. The Chemistry of Soils. Oxford University
Press, New York.
Stiff, H.A. and L.E. Davis. 1952. A Method for Predicting the
Tendency of Oil Field Water to Deposit Calcium Carbon-
ate. AIME Trans. Petroleum Div. 195:213-216
Stumm, W. and JJ. Morgan. 1981. Aquatic Chemistry, 2nd
ed. Wiley Interscience, New York.
Tabak, H.H. et al. 1981. Biodegradability Studies with Or-
ganic Priority Pollutant Compounds. J. Water Pollution
Control Federation 53(10): 1503-1518.
Tenorio, P. A., R.H.F. Young, and H.C. Whitehead. 1969.
Identification of Return Irrigation Water in the Subsur-
face. Water Resources Research Center Tech. Report 30,
University of Hawaii, Honolulu, HI.
Thornton, I. (ed.). 1983. Applied Environmental Gochemis-
try. Academic Press, New York.
Thorstenson, D.C., D.W. Fisher, and M.G. Croft. 1979. The
Geochemistry of the Fox Hills-Basal Hell Creek Aquifer
in Southwestern North Dakota and Northwestern South
Dakota. Water Resources Research 15(6):1479-1498.
Thurman, E.M. 1985. Humic Substances in Groundwater. In:
Humic Substances in Soil, Sediment, and Waten Geo-
chemistry, Isolation, and Characterization, Aiken, G.R.,
D.M. McKnight R.L. Wershaw, and P. MacCarthy (eds.),
John Wiley & Sons, New York, pp. 87-103.
Tryon, C.P. 1976. Ground-Water Quality Variations in Phelps
County, Missouri. Ground Water 14(4):214-223.
U.S. Environmental Protection Agency (EPA). 1989. Assess-
ing the Geochemical Fate of Deep-Well-Injected Hazard-
ous Waste A Reference Guide. EPA/625/6-89-025a.
Van Beek, C.G.E.M. and J. Van Puffelen. 1988. Changes in
the Chemical Composition of Drinking Water after Well
Infiltration in an Unconsolidated Sandy Aquifer. Water
Resources Research 23(l):69-76.
Van Cleemput, 0. and A.S. El-Sebaay. 1985. Gaseous Hydro-
carbons in Soil. Advances in Agronomy 38:159-181.
White, R.E. 1985. The Influence of MacroPores on the Trans-
port of Dissolved and Suspended Matter Through Soil.
In: Advances in Soil Science, B.A. Stewart (ed.), Springer-
Verlag, New York, Vol. 3.
White, D.E., J.D. Hem, and G.A. Waring. 1963. Chemical
Composition of Subsurface Waters. U.S. Geological Sur-
vey Professional Paper 440-F.
Wilson, J.T. and J.F. McNabb. 1983. Biological Transforma-
tion of Organic Pollutants in Groundwater. Eos (Trans.
Am. Geophysical Union) 20:997-1002.
Wilson, L.C. and J.V. Rouse. 1983. Variations in Water
Quality During Initial Pumping of Monitoring Wells.
Ground Water Monitoring Review 3(1): 103-109.
Wood, W.W. and R.L. Bassett. 1973. Chemical Quality of
Recharge Water as a Function of Bacterial Activity Be-
neath a Recharge Basin. Eos (Trans. Am. Geophysical
Union) 54:261.
Wood, P.R., R.F. Lang, and I.L. Payan. 1985. Anaerobic
Transformation, Transport and Removal of Volatile Chlo-
rinated Organics in Ground Water. In: Ground Water
Quality, C.H. Ward, W. Giger, and P.L. McCarty, (eds.),
Wiley Interscience, New York, pp. 493-511.
Yamn, B., G. Dagan, and J. Goldschmid (eds.). 1984. Pollut-
ants in Porous Media: The Unsaturated Zone Between
Soil Surface and Groundwater. Springer-Verlag, New
York.
Zimmie, T.F. and C.O. Riggs (eds.) 1979. Permeability and
Groundwater Contaminant Transport. ASTM STP 746.
American Society for Testing and Materials, Philadel-
phia, PA.
Zehnder, A.J.B. (ed.). 1988. Biology of Anaerobic Microor-
ganisms. Wiley-Interscience, New York.
Zehnder, AJ.B. and W. Stumm. 1988. Geochemistry and
Biogeochemistry of Anaerobic Habitats. In: Biology of
Anaerobic Microorganisms, A.J.B. Zehnder (ed.), Wiley-
Interscience, New York, pp. 1-38.
ZoBell, C.E. 1946. Studies on Redox Potential of Marine
Sediments. Bull. Am. Ass. Petroleum Geol. 30(4):477-
513.
122
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Chapter 9
Geochemical Sampling of Subsurface Solids and Ground Water
J. Russell Boulding and Michael J. Barcelona
9.1 General Considerations
9.1.1 Types of Monitoring
A complete sampling program for subsurface site charac-
terization includes several types of monitoring, each with its
own goal. The goal of detection monitoring is generally to
determine the presence of contaminated conditions. Unfortu-
nately, drinking water wells have been among the most com-
mon detective monitoring systems historically. Assessment
monitoring seeks to identify the extent and magnitude of
contamination. If assessment monitoring results indicate a
degree of contamination requiring remediation, evaluation
monitoring is used to provide data necessary to design the
remediation system. Performance monitoring is designed to
evaluate the success of remediation efforts. Each stage of
monitoring often requires the placement of additional moni-
toring wells and piezometers for water level measurements.
Other types of monitoring include litigation monitoring in
response to legal actions at contaminated sites and research
monitoring aimed at specific scientific objectives.
9.1.2 Sampling Protocol
The goal of a sampling program with an overall monitor-
ing design is often to avoid underestimating a particular
impact either in terms of concentration or spatial distribution.
Characterization of geochemical variability also is necessary
to identify potential chemical problems that may affect selec-
tion and design of systems for ground-water treatment.
The field sampling protocol is often the weakest link in
soil and ground-water sampling programs. Most initial effort
and fiscal resources should be spent on characterizing basic
site geology and hydrology. An optimal program may call for
the placement of three or four times as many piezometers than
wells for water-quality sampling. Initial selection of locations
for sampling must be based on a good preliminary character-
ization of the geology and hydrogeology of the site. This may
require spending more of the available financial resources on
hydrogeologic characterization than on chemical sampling
and analyses. Additional sample locations should be added as
understanding of the site evolves.
As discussed in Section 7.1.4, sample location and fre-
quency are among the most critical aspects of sampling be-
cause sample collection and sample analysis sometimes can
give entirely erroneous results even when approached and
executed carefully. Good vertical and horizontal resolution of
hydrogeologic conditions are essential before choosing sample
locations. Uncertainty, hydrogeologic variability, and quality-
assurance decision-making need to be addressed from the
initial design stage. Later, an effective sampling strategy and
written protocols should be prepared. These measures can
improve confidence in subsequent chemical results. Docu-
mentation of all sampling procedures is essential, because
data collected for a particular purpose may end up being used
and interpreted for other objectives.
Sampling protocols should leave room for evolutionary
development of the network design. For example, sampling
experiments can be used to determine spatial correlation for
solid samples. A large number of surface samples or split-
spoon samples can be collected but it may only be necessary
to analyze a certain percentage (20 to 50 percent) to achieve
adequate spatial coverage. If the initial sample groups indicate
sufficient sampling resolution, the other samples need not be
analyzed. If necessary, additional samples can be analyzed
until geostatistical analysis indicates an adequate sampling
intensity has been achieved. Samples should not be thrown
away if there is any possibility that somebody may use them
in the future and if adequate preservative measures are fea-
sible.
Many references thoroughly cover one or more aspects of
developing a sampling program and protocol for subsurface
solids and/or ground-water. Table 9-1 lists and summarizes 29
of these major reference sources. Rehm et al. (1985) probably
contains the most comprehensive review of the literature on
methods for hydrogeologic investigations up to 1985. The rest
of this chapter focuses on developments since that time,
although particularly relevant pre-1985 references are occa-
sionally cited.
Table 9-2 lists sources of information on four aspects of
general sample design: (1) general theory, (2) soil sampling,
(3) vadose zone sampling, and (4) ground-water sampling.
General aspects of selecting sample location, frequency, and
size are discussed in the remainder of this section. Section 9.2
reviews sampling of subsurface solids and vadose zone water
further, and Section 9.3 covers sampling of ground water.
123
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Table 9-1. Major Reference Sources on Subsurface Sampling Methods
Reference
Description
Barcelona et at. (1983)
Barcelona et al.
(1985a)
Berg (1982)
Collins and Johnson
(1988)
Dunlapetal. (1977)
Everett et al. (1976)
Fennetal. (1977)
Ford etal. (1984)
Geo Trans (1989)
Gibb etal. (1981)
Gillham etal. (1983)
Holden (1984)
Howsam (1990)
NCASI(1982)
Nielsen (1991)
NJDEP(1988)
Niaki and Broscious
(1986)
Oudjik and Mujica
(1989)
Rehm etal. (1985)
Scalfetal. (1981)
Simmons (1991)
Sisk(1981)
Summers and Gherini
(1987)
Tinlin (1976)
Toddetal. (1976)
UNESCO (1983)
U.S. EPA (1985)
U.S. EPA (1986a)
U.S. EPA (1986b)
U.S. Geological
Survey (1977+)
van Duijvenbooden and
van Waegeningh (1987)
Wood (1976)
Guide to selection of materials for monitoring well construction and ground-water sampling.
Ground-water sampling guide covering QA/QC procedures, analyte selection, drilling methods,
monitoring well design, well development, sampling, and recommended sampling protocols.
Handbook focusing on all aspects of water and wastewater sampling, sample preservation, and QA/QC
procedures. One chapter covers sampling of ground water and another covers sampling/preservation and
storage considerations for trace organics.
Contains 37 papers on field methods on the investigation of ground-water contamination.
EPA report on sampling for organic chemicals and microorganisms in the subsurface.
EPA report describing ground-water-related measuring techniques applicable to the land surface, topsoil, vadose
zone and zone of saturation. Also presents cost data on various methods.
Procedures manual for ground-water monitoring at solid waste disposal facilities. Covers monitoring networks,
monitoring and well technology, chemical parameters for indicators ofleachate and sampling.
Manual covering sampling methods for solids, gases, and liquids at hazardous waste sites.
Manual developed for the electric utility industry detailing the design, implementation, and maintenance of a
ground-water monitoring program.
Contains recommendations for procedures to collect representative ground-water quality samples based on tests
of different procedures at six monitoring wells at waste disposal sites in Illinois.
Focuses on sources of sample bias resulting from hydrogeologic factors and chemical alterations; examines
chemical characteristics of inorganic and organic parameters, sampling installations, sample collections and
methods.
Primer focusing on ground-water sampling for volatile organic compounds.
Proceedings of the international conference on water well monitoring, maintenance, and rehabilitation.
Guide to ground-water sampling with chapters on preparation for sampling, sample collection, and sample
pretreatment and field analysis.
Handbook covering all aspects of vadose zone and ground-water monitoring.
Manual on field sampling procedures prepared by New Jersey Department of Environmental Protection.
EPA report describing over 30 methods for detecting leaks in underground tanks.
Handbook focusing on field methods for identification, location, and investigation of pollution sources affecting
ground water.
Comprehensive review of methods for solids, unsaturated zone, and ground-water physical and chemical
characterization. Bibliography contains over 600 references on these topics.
Manual covering drilling methods, collection of ground-water samples, field tests and preservation, with a short
chapter on sampling subsurface solids.
Edited volume covering sampling and analysis of hazardous wastes.
Manual for ground-water/subsurface investigations at hazardous waste sites. Appendix on information sources is
especially useful.
Manual focusing on water sample QA/QC procedures, and procedures for collecting samples.
EPA report containing nine case studies illustrating procedures for monitoring various classes of ground-water
pollution sources.
EPA report describing a 15-step monitoring methodology for ground-water quality.
Proceedings of an international symposium on methods and instrumentation for the investigation of ground water;
contains over 60 papers.
EPA guide for developing administrative orders to address RCRA ground-water monitoring violations at interim-
status land disposal facilities.
EPA's RCRA ground-water monitoring technical enforcement guidance document.
EPA test methods for evaluating solid waste. Part IVin Volume II (Field Methods) defines acceptable and
unacceptable designs and practices for ground-water monitoring.
USGS National Handbook of Recommended Methods for Water Data Acquisition. Individual
chapters have come out at different dates. Pertinent chapters include: (2) Ground Water (1980); (4) Biological
and Microbiological Quality of Water (1983); (5) Chemical Quality (1982); and (6) Soil Water (1982).
Proceedings of international conference containing a number of papers on soil and ground-water monitoring
strategies and vulnerability mapping.
USGS guidelines for collection and field analysis of ground-water samples for selected unstable constituents.
124
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Table 9-2. Sources of Information on General Sample
Design
Topic References
General theory
Elementary Gilbert (1987), Slonim (1957), Tanur
(1978), Williams (1978)
Advanced Cochran (1977), Deming (1950), Hansen
etal. (1953), Hendricks (1956),
Jesse/7 (1978), Kish (1965), Pitard
(1989), Sukhatme and Sukhatme
(1970), Yates (1981)
Table 9-3. Summary of Sampling Designs and Conditions for
Their Use
Soil
Sample
Design
Compositing
Vadose Zone
Monitoring
Concepts
Network
Design
Ground water
Compositing
General
Network
Design
Sample
frequency
Bart/? etal. (1989), Dietal. (1989),
Hoffman (1986), Loehr et al (1986),
Peterson and Calvin (1986), Williams
etal. (1989). See also Table 7-7
Peterson and Calvin (1986),
Williams etal. (1989)
Everett etal. (1982), Everett etal.
(1984), Kirschner and Bloomsburg
(1988)
Bumbetal. (1988), McKeeandBumb
(1988)
Rajagopal and Williams (1989)
Steele (1986)
Hsueh and Rajagopal (1988), Hughes
and Lettenmaierf 1981), Loaiciga
(1989), McNichols and Davis (1988),
Nightingale and Bianchi (1979),
Sophocleous etal. (1982)
Close (1989), Hsueh and Rajagopal
(1988), Loaiciga (1989), Sgambatand
Stedinger (1981), Rajagopal (1986)
9.1.3 Sample Location
Table 9-3 summarizes major types of sampling designs
and when they should be used for characterizing subsurface
geochemistry. In general, haphazard water-quality or solid
sampling is not an appropriate approach to designing sam-
pling for subsurface geochemical characterization, even though
professional judgment alone, is probably the most frequently
used method for siting ground-water monitoring wells. Figure
9-1 illustrates some two-dimensional probability sampling
designs for spatial characterization. The trends or patterns that
commonly exist in subsurface contamination mean that simple
random sampling will not give as accurate an estimate of
population characteristics as stratified random and grid sam-
pling designs.
Hydrogeologic characterization, initially using surface
geophysical techniques followed by piezometers and prelimi-
nary well tests to estimate the distribution of hydrogeologic
parameters, should come before the location and installation
of monitor wells, Good vertical resolution is essential in
sampling to characterize distribution of oxidized and reduced
species, contaminants, and microbiota. Achieving this resolu-
tion requires more discrete well completions with short screens.
Type of Sampling
Design
Haphazard
sampling
Judgment
sampling
Probability
sampling
Simple
random
Stratified
random
Multistage
Cluster
Systematic
Double
Search
Sampling
Conditions When the Sampling
Design is Useful
A very homogeneous population over
time and space is essential if unbiased
estimates of population parameters
are needed. This method of selection
is not recommended due to difficulty in
verifying this assumption.
The target population should be clearly
defined, homogeneous, and
completely assessable so that sample
selection bias is not a problem. Or
specific environmental samples are
selected for their unique value and
interest rather than for making
inferences to a wider population.
The simplest random sampling design.
Other designs below will frequently
give more accurate estimates of
means in the population that contains
trends or patterns of contamination.
Useful when a heterogeneous population
can be broken down into pans that
are internally homogeneous.
Needed when measurements are made
on subsamples or aliquots of the field
sample.
Useful when population units cluster
together (schools offish, clumps of
plants, etc.) and every unit in randomly
selected clusters can be measured.
Soil and ground-water contamination
rarely, if ever, exhibit this
characteristic.
Usually the method of choice when
estimating trends or patterns of
contamination over space. Also useful
for estimating the mean when trends
and patterns in concentrations are not
present or they are known a priori or
when strictly random methods are
impractical.
Useful when there is a strong linear
relationship between the variable of
interest and a less expensive or more
easily measured variable.
Useful when historical information, site
knowledge, or prior samples indicate
where the object of the search may be
found.
Source: Adapted from Gilbert (1987)
In most cases, 5-ft to 1.5-m well screens should give adequate
vertical resolution.
The spatial distribution of contamination is a major con-
cern with sampling solids. The intensity and number of samples
depends on the nonsampling variance, which is the variability
of concentration that is unrelated to sampling procedures.
Spatial structure determines the distance between samples
that have essentially the same concentration, called the range
of correlation, to avoid oversarnpling (see Section 7.3.2).
125
-------
Simple Random
Sampling
Stratified Random
Sampling
(a)
•
• .
.* .
•
•
•
•
• •
••
(c)
Strata
Two-Stage
Sampling
/
Pr
Un
x
mary
its
.'
— |
**
Cluster
Sampling
(d)
Clusters
Systematic Grid
Sampling
ffl
Random Sampling
Within Blocks
•
•
•
•
«
•
•
•
•
•
•
•
Figure 9-2. Palmerton wind rose, 1978-1979 data (from Starks
et al., 1986).
Figure 9-1. Some two-dimensional probability sampling designs
for sampling over space (from Gilbert, 1987). See
Table 9-3 for description of when these sampling
designs are useful.
There are two broad designs for soil sampling: (1) grids
in which samples are taken from a matrix of squares or
quadrants at a site, and (2) transects in which samples are
taken at specified intervals along a line. Figure 7-10 in Chap-
ter 7 shows contours of lead concentration in soil drawn from
grid sampling. Grids presume an aerial or dispersed source of
some kind, and transects presume a preferential source. For
example, Starks et al. (1986) established sampling transects
where the length was proportional to the frequency with
which wind blew in a particular direction to characterize
metal contamination from a smelter near Palmerton, Pennsyl-
vania (see Figures 9-2 and 9-3). Flatman (1986) describes use
of geostatistics for determining sampling intensity. Grids can
be used to estimate short-range correlation. Transects along
the path of ground-water or contaminant movement provide
the best way to look at long-range correlation. The combina-
tion of the two strategies coupled with the initial analysis of
selected solid samples at alternate grid or transect locations
can be quite effective.
The combined strategy also can avoid the potential col-
lection of redundant information. Using geostatistical analysis
techniques of successive analytical subsets minimizes the
number of samples actually analyzed. Transects could be both
parallel and perpendicular to the axis of ground-water move-
ment, along with some random samples from a grid, as shown
in Figure 9-1 (f). Analysis of samples from four equally spaced
locations on a transect or grid within the area of influence is a
good starting point to estimate the distance of short-range
correlation. For soils, at least 5 percent of sampling points
should be duplicated to help determine the sampling variabil-
ity, so it can be analyzed with geostatistical techniques. At
least 5 percent of the samples should be split as well.
Preliminary efforts that can help guide the location of
initial wells for ground-water sampling include (1) surface
geophysical techniques for mapping extent of contaminant
plumes; (2) soil gas sampling techniques; (3) Hydropunch®
sampling; and (4) selective sampling of piezometers for simple
constituents such as pH, conductance, and possibly iron or
dissolved oxygen concentrations.
Soil gas monitoring (see Section 5.2.5) and Hydropunch®
ground-water sampling (see Section 9.3.4) probably give the
best pictures of short-range variability in three dimensions.
Sampling from monitoring wells usually gives some sort of
integrated value depending on the relative width or thickness
of the hydrogeologic formation of interest and the length of
the screen. Disadvantages of soil gas concentrations include
(1) lack of the ability to directly calibrate, because all values
are relative and difficult to reproduce, (2) decontamination,
and (3) short circuiting of air from the surface, which can
distort results.
126
-------
To S Miles
1200'Points
N
Figure 9-3. Sample pattern for the initial Palmerton survey (1"= 4250') (from Starks et al., 1986).
9.1.4 Sampling Frequency
Table 9-4 shows estimated ranges of sampling frequency
in months necessary to maintain information loss at less than
10 percent for selected types of chemical parameters. For
many chemical constituents, quarterly sampling is adequate
for characterizing short-term (i.e., monthly to 1 or 2 years)
changes over time. For some reactive constituents such as iron
and other redox-sensitive constituents, bimonthly sampling
may be required.
With intermittent sources of contamination, it is espe-
cially important that the frequency of sampling not allow a
contaminant to be missed. Barcelona et al. (1985a) describe a
procedure for estimating sample frequency to detect contami-
nant plumes based on the type of plume (slug, intermittent, or
continuous) and hydrogeologic parameters of gradient, hy-
draulic conductivity, effective porosity, and distance along
the flow path. Figure 9-4 shows a nomograph that can be used
when these parameters are known. When the contaminant
plume is a slug source or intermittent, sampling frequency
should probably be more frequent to ensure that the plume is
not missed. One advantage to the slow movement of ground
water is that if there are questions about a sample, resampling,
a week later will yield roughly the same ground water.
Precise estimation of optimum sampling frequency is
probably impractical for most investigations. For example,
Bell and DeLong (1988) found that tetrachloroethylene at
concentrations of 200 to 300 ug/L exhibited variations of a
factor of one or two over the course of a year. Their work
points out that data collection may be required for 4 years or
more in order to estimate the optimal sampling frequency to
determine seasonal variability. Therefore, it is important to
select sampling frequency on the basis of an initial period of
monitoring in the context of the duration of the program.
It should not be necessary to sample all monitoring wells
every time samples are taken. Sampling selected wells can
develop a preliminary picture, with additional follow-up sam-
pling at additional wells rounding out the picture.
127
-------
9.1.5 Sample Type and Size
Soil sampling must take into account fractures in earth
materials and the fact that the subsurface is heterogeneous (as
scales ranging from centimeters to meters). If the soil has
obvious fractures and channels in the subsurface, sampling
should sample both affected and apparently nonfractured ar-
eas for comparison. Soil sample quantities of less than 100 g
tend to be unrepresentative even of the areas where the sample
is taken. In the laboratory, the sample can be mixed, subsampled
prior to analysis.
Compositing samples is often beneficial for soil investi-
gations. However, where volatile constituents are involved,
compositing is not practical because handling samples in the
air for compositing will result in the loss of the contaminant.
One way to get around this problem is to take two or three
samples within each identifiable core segment and put them
into a sealed glass vial immediately after sampling. In this
case, a volume of methyl alcohol in the sealed vial can
improve volatile recovery and expedite analysis. However, it
is possible that the sampling variance from potential loss of
Table 9-4. Estimated Ranges of Sampling Frequency (in
Months) to Maintain information Loss at <10% for
Selected Types of Chemical Parameters
Pristine
background
Type of Parameter conditions
Contaminated
Upgradient Downgradient
Water quality
Trace constituents 2 to 7 1 to 2 2 to 10
(<1.0 mg/L)
Major constituents 2 to 7 2 to 38 2 to 10
Geochemical
Trace constituents 1 to 2 <2 1 to 5
(<1.0mg/L)
Major constituents 1 to 2 7 to 14 1 to 5
Contaminant indicator
TOO
TOX
Conductivity
PH
2
6 to 7
6 to 7
2
3
24
24
2
3
7
7
1
10
(F)
10'
10'
10><
10'
10 ^
^^^
.1
(D)
WO—,
80
60
40
20
10 —
8
6
4
2
1.0 _
.8 •
.6 .
1
¥
I
£
g Flow Pa
^
u
§
i.
^
(N)
.05 -i
.08 •
•10 • &
'OS
•»5-|
.20 - .|
.25 • |
.30^r Uj
^<40 -
' .50 •
ON
Example (clean sand)
K= 10'
1= 104
N = 0.30
D = 0.4 meters
Figure 9-4. Sampling frequency nomograph (from Barcelona et al., 1985).
128
-------
volatiles involved in handling the sample may far exceed the
actual variability in the field.
Williams (1989) compares the results of one 500-g sample,
twenty 25-g composite samples, and ten 50-g composite
samples. He found that a single 25-g composite sample was
the most accurate and precise technique for determining ra-
dium concentrations in contaminated surface soil. Initial soil
samples of 100-g are about the best size for such composite
analyses.
9.1.6 Vadose versus Saturated Zone
Careful sampling of gases and solids in the vadose zone
can provide information for better locating monitoring wells
in the saturated zone. The mass of contaminant, or at least the
most persistent contaminants, are often associated with the
solids.
9.2 Sampling Subsurface Solids and Vadose
Zone Water
9.2.1 Analyte Selection
Halocarbons, chlorinated hydrocarbon solvents (e.g., tetra-
and trichloroethylene), and fuel constituents (e.g., toluene,
benzene, ethyl benzene, and xylenes) are amenable to prelimi-
nary delineation by soil gas methods. Soil gas samples for
carbon dioxide, methane, oxygen, and nitrogen can provide
additional insights into subsurface chemistry, particularly mi-
crobiological activity.
In addition to examining chemical constituents, analyz-
ing solid samples for grain-size distribution and correlation
with permeability can be helpful.
9.2.2 Sampling Devices and Techniques
Table 9-5 lists sources of information on sampling soil
and vadose zone solids, water, and gases.
Simple techniques for surface sampling of soils include
the hand auger, brace and bit, and posthole diggers. The most
commonly used core sampling devices are split spoons or
Shelby tubes that provide a continuous or driven core during
drilling operations. Sampling continuously or ahead of hol-
low-stem- drilling augers are good ways to obtain uncontami-
nated and minimally disturbed soil samples. Section 3.1
provides some additional discussion of these sampling meth-
ods for obtaining information on subsurface stratigraphy.
Where the surface layer of soil is known to be heavily
contaminated, as with sites involving smelters and uranium
mills, the surface should be scraped away before sampling at
lower levels so the sampler is not contaminated as it passes
through the contaminated surface.
Figure 9-5 shows a soil core sampling apparatus de-
scribed by Myers et al. (1989) that can obtain undisturbed
cores for laboratory leaching experiments. A variety of sam-
plers are available that advance in front of an auger. The better
devices have a plunger or cylinder that maintains a partial
vacuum to prevent the soil material from falling out of the
core (Munch and Killey, 1985; and Zapico et al., 1987). This
vacuum is particularly important for saturated sands that
simply flow out of normal sampling tubes. Figure 9-6 shows a
modified wireline piston design for sampling cohesionless
sediments and Figure 9-7 shows how this device can be used
to take samples through a hollow-stem auger. In careful use,
the more sophisticated devices can achieve a 50 percent core
recovery. Heaving sands create special problems. Filling the
auger with water sometimes helps prevent clogging from
heaving sands by maintaining hydrostatic pressure.
Suction lysimeters can be used to sample pore water in
the vadose zone. Extremely variable transmissive properties
of surface soils make accurate interpretation of soil pm water
concentrations very difficult. Virtually all of the water move-
ment and associated contaminant transport may occur in about
5 percent of the soil profile. The zone of sampling influence
with a suction lysimeter is about 10 or 20 cm for a 24-hour
period (Morrison and Lowery, 1990). In some instances,
longer suction sampling periods may extend the influence to
50cm.
Sampling for microbiological parameters requires both
the collection of soil samples and the paring of any outside
portion that may have been in contact with the sampling
apparatus. This operation should be done before placing the
samples in sterile glass vials.
Table 9-5. Sources of Information on Sampling Soil and
Vadose Zone Solids, Solutes, and Gases
Topic References
Cohesive
Noncohesive
Volatiles in Soil
Vadose zone solute
sampling methods
Overviews
Physical Properties
Moisture Potential
Moisture Content
Solute Sampling
Soil gas sampling
Gas Properties
Overviews
Case Studies
Barth and Starks (1985), Cameron et al.
(1966). Fordetal. (1984), Rehm etal.
(1985), Mason (1983), Myers etal.
(1989)
Munch and Killey (1985), Zapico et al.
(1987), Armstrong etal. (1988)
Slater and McLaren (1983)
Everett etal. (1982, 1983, 1984), Rehm
etal. (1985), U.S. EPA (1986c),
Wilson (1980), Wilson (1983)
Wilson (1982)
Wilson (1981)
See overview references.
Brown (1987), Everett and McMillion
(1985), Johnson and Cartwright
(1980), LJtaor (1988), Stevenson
(1978)
Mackay and Shiu (1981)
Devittetal. (1987), Kerfoot and Barrows
(1986), Marrin (1987), Marrin and
Kerfoot (1988)
See Table 9-6
129
-------
Pusher with Depth
Adjustment
PVC Coupler
Spacer Ring
Soil Core Cutter
120 cm.
Figure 9-5. Undisturbed soil core sampling apparatus (from Myers et al., 1989).
Soil gas sampling generally involves driving a probe into
the subsurface. Typically, the probes are driven by hand or
with some kind of pneumatic or electric hammer. Soil gas is
obtained by applying a vacuum that brings the soil gas into the
vicinity of the tip of the probe. Samples are collected in
fluorocarbon bags or syringes and analyzed on site or in a
laboratory. Analysis techniques can be as simple and nonspe-
cific as a hand-held gas survey meter, and as detailed and
specific as an analytic laboratory's instrumentation allows.
Mobile laboratories provide an intermediate level of analyti-
cal detail; they provide semiquantitative results with precision
on the order of plus or minus 100 percent. At least 5 percent of
air-filled porosity is required to pull a vacuum to obtain
samples.
Table 9-6 summarizes information from 14 soil gas in-
vestigations. Soil gas samples for areal characterization are
usually taken at a uniform depth with the specific depth
typically from 1 to 6 ft below the surface, although Glaccum
et al. (1983) sampled immediately above the water table,
which was as much as 10 m deep. Vertical profiles may
provide additional insight into contaminant behavior. Figure
9-8 shows six types of vertical concentration profiles that
develop under different subsurface conditions. Special care
should be taken to identify any underground utility lines to
avoid accidental puncture with the probe. Buried sewers or
product lines may be the source of soil gas contamination, and
other utility lines may provide a directional component to
contamination (Marrin and Thompson, 1987). See Section
5.2.5 for additional discussion of soil gas sampling methods.
9.3 Sampling Ground Water
Figure 9-9 shows a generalized flow diagram of ground-
water sampling steps, and Table 9-7 lists additional sources of
information on various aspects of ground-water sampling.
130
-------
1
2
3
4
Teflon Wiper Disc
Brass Bushing
Neoprene Seals
Swivel
Figure 9-8. Modified wireline piston design (from Armstrong et al., 1988).
References that provide good general coverage of ground-
water sampling include Barcelona et al. (1985), GeoTrans
(1989), Gillham et al. (1983), Rehm et al. (1985), and Scalf et
al. (1981).
9.3.1 Analyte Selection
Tables 9-8 and 9-9 identify chemical constituents of
interest for various types of ground-water monitoring activi-
ties. In hazardous waste site investigations, regulations will
generally specify the contaminants to be tested for. Focusing
on priority pollutants alone, however, may not provide a
complete geochemical picture of contamination. The source
of contamination may involve a large number of individual
contaminants that are not classified as hazardous. Also, deter-
mination of redox-sensitive constituents (dissolved oxygen
and dissolved iron), pH, and conductance, may provide valu-
able insight into subsurface contaminant geochemistry.
Highly mineralized ground water, commonly encoun-
tered in formations being evaluated for deep-well injection of
wastes, may require more complete analyses for natural inor-
ganic and organic constituents. Table 9-10 lists analytical
results for ground-water samples from four deep-well injec-
tion sites and the Frio formation in Texas (which has received
more deep-well injected wastes than any other formation in
the United States). Not all of the studies analyzed the same
constituents in all samples, but an examination of this table
may give some guidance for analyte selection.
Iron, an inexpensive constituent to determine analyti-
cally, can be used as an indicator of redox conditions and
potential mobility for heavy metals. Dissolved gases are ex-
cellent indicators of redox conditions and microbial activity.
For example, Leenheer and Malcolm (1973) analyzed for H2
N2, CH4, CO2, and H.S in serial samples from a well through
which a plume of deep-well injected wastes passed. They used
changes in the relative percentages of the different gases as
indicators of changing microbial activity. Malcolm and
Leenheer (1973) suggest that separate analysis for dissolved
organic carbon DOC) and suspended organic carbon (SOC)
can yield more complete analytical results.
Calcium carbonate and iron/manganese concentrations
are especially important parameters if remediation involves
air stripping. Air-stripping towers are particularly susceptible
to fouling by calcium carbonate and metal oxide precipitates.
131
-------
o
Wireline
Hammer
Borehole
Drill Rods
"Hollow-Stem
Augers
Core Barrel
Drill Bit
Advancement
(Sampling)
Placement
Drill Rods
Wireline
Borehole
Drilling Mud
Filter Cake
Hollow-Stem Auger
Core Barrel
Liner
X!Lx- Piston
Drill Bit
Aquifer
Wireline Recovery
1
/.
J
i
\r
^
i
»
u
.
m
p
•
.
«4
0
1
r
Sample
D/vY/B/f -
Drilling Mud
Sample Hole
Figure 9-7. Wireline piston core barrel sampling operation (from Zapico et al., 1987).
When trichloroethylene (TCE) is involved as a contami-
nant, it is important also to analyze for biotransformation
products (e.g., 1,2-dichloroethene, 1,1-dichloroethene, and
vinyl chloride), The vinyl chloride monomer metabolic prod-
uct is more toxic than TCE and resistant to degradation under
anaerobic conditions.
Battista and Connelly (1989) found that inorganic param-
eters such as chemical oxygen demand, specific conductance,
chloride, alkalinity, and hardness were reasonably good indi-
cators for predicting VOC contamination from landfills. When
the inorganic parameters were detected above background
levels in monitoring wells, VOCs were also usually present.
Out of 49 ground-water samples at landfill sites in Wisconsin,
VOCs and elevated inorganic parameters were detected at
about the same frequency in 20 (41 percent), elevated inor-
ganic parameters without VOCs were detected in 11 (22
percent), and VOCs without elevated inorganic parameters
were detected in 3 wells (6 percent). The remaining 15 wells
in the study showed neither VOCs nor elevated inorganic
parameters.
9.3.2 Well Development
Well development, which involves the removal of fines
created during the drilling process, is essential before sam-
pling begins. Pumping rates generally used for well develop-
ment are 5 to 10 gpm. Bailing, swabbing, pumping, and
air-lifting are common methods used for development. Table
4-2 compares the advantages and disadvantages of the most
commonly used well development techniques. Air develop-
ment may increase the possibility of environmental exposure
to workers at the surface where volatiles are involved.
132
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Table 9-6. Soil Gas Sampling Case Studies
Location
Contaminant/Soil Gas Sample Methods
Reference
Las Vegas, NV
Tucson, AZ
(Water table 120')
Denver, CO
Northern CA
(Water table 25')
New England
Sudbury, MA
Battle Creek, Ml
Not specified
Sandwich, MA
(Water table 7')
4 unspecified
locations
Military facility
2 unspecified sites
Benzene, chlorobenzene. Dynamic samples above water table
(up to 10 m deep).
Trichloroethylene (TCE). Dynamic, area/ (<2 m); vertical (6 m intervals).
Tatrachloroethylene (PCE). 3-day static samples near surface.
1,1,2-Trifluorotrichloroethane (F-113). Dynamic, area (10'), vertical to 20'.
TCE. Dynamic, area/ at 18".
Gasoline. Dynamic, area/ at 18 and 30" (Site 1), 12 and 24" (Site 2).
1,2-Dichloroethene, PCE. Dynamic, arealat4.5:
Gasoline. Dynamic, area/ (depth not specified).
Gasoline. Dynamic, multiple vertical profiles atlft intervals to 8'.
Volatile hydrocarbons (3 sites); TCE (1 site). Dynamic, area! at 3'.
Diesel fuel. Dynamic, vertical profiles at 2-3' intervals to around 10'
Chlorinated solvents (TCE, TCA, vinyl chloride, F-113. Dynamic,
areal at 4 to 6',
Glaccum et al. (1983)
Martin and Thompson (1984,
1987)
Voorhees et al. (1984)
Martin and Thompson (1984)
Spittler etal. (1985)
Spittler et al. (1985)
Wittmann etal., (1985)
Goodwin and Burger (1989)
Kerfoot and Soderberg (1988)
Newman etal. (1988)
Diem etal. (1988)
Shangraw et al (1988)
(A)
MW&
Depth
VOC
Concentration
(B)
Depth
VOC
Concentration
Depth
VOC
Concentration
Depth
(E)
Depth
VOC
Concentration
(A) Homogeneous Porous Material with Sufficient Air-filled Porosity
(B) Impermeable Subsurface Layer (e.g., Clay or Perched Water)
(C) Impermeable Surface Layer (e.g., Pavement)
(D) Zone of High Microbiological Activity (Circles and Wavy Lines Indicate Different Compounds)
(E) VOC Source in the Vadose Zone
Figure 9-8. Soil-gas ooncentrations under a variety of conditions (from Marrin and Kerfoot, 1988).
133
-------
Step
Well Inspection
Procedure
Hydrologic Measurements
I
Wai! Purging Removal or Isolation of Stagnant Water
I
Determination of Well-Purging Parameters
(pH, Eh. T. £1-')"
Sample Collection
Filtration' Unfiltered
Field I - ' -
Determinations" \
Volatile Organics, TOX
Dissolved Gases, TOC
Large Volume Sam-
ples for Organic
Compound Determi-
nations
Field Filtered"
Preservation
Field Blanks
Standards
Storage
Transport
Assorted Sensitive
Inorganic Species
NO,, A/H/, Fe(ll)
(as needed for good
OA/OC)
Trace Metals for
Mobile Substance
Load+++
Alkalinity/Acidity"
i
Trace Metal Samples
for Specific Geochemical
lnformation+++
S', Sensitive
Inorganics
Major Cations and
Anions
Essential Elements
Water-Level
Measurements
Representative Water
Access
Verification of
Representative Water
Sample Access
Sample Collection by
Appropriate Mechanism
Minimal Sample Handling
Head-Space
Free Samples
Minimal Aeration or
Depressurization
Minimal Air Contact,
Field Determination
Adequate Rinsing against
Contamination
Minimal Air Contact,
Preservation
Minimal Loss of Sample
Integrity Prior to Analysis
Recommendations
Measure the water level to ±0.3
cm (±0.01 ft).
Pump water until well purging
parameters (e.g., pH, T, Q-', Eh)
stabilize to ±10% over at least
two successive well volumes
pumped.
Pumping rates should be limited
to ~ 100 mL/min for volatile
organics and gas-sensitive
parameters.
Filter: Trace metals, inorganic
anions/cations, alkalinity.
Do not filter: TOC, TOX, volatile
organic compound samples. Filter
other organic compound
samples only when required.
Samples for determinations of
gases, alkalinity andpH should
be analyzed in the field if at all
possible.
At least one blank and one
standard for each sensitive
parameter should be made up in
the field on each day of
sampling. Spiked samples are
also recommended for good QA/
QC.
Observe maximum sample
holding or storage periods
recommended by the Agency.
Documentation of actual holding
periods should be carefully
performed.
Denotes samples that should be filtered to determine dissolvad constituents. Filtration should be accomplished preferably with in-line filters
and pump pressure or by N2pressure methods. Samples for dissolved gases or volatile organics should not be filtered, in instances where
well development procedures do not allow for turbidity free samples and may bias analytical results, split samples should be spiked with
standards before filtration. Both spiked samples and regular samples should be analyzed to determine recoveries from both types of handling.
Denotes analytical determinations that should be made in the field.
See Puls and Barcelona (1989).
Figure 9-9. Generalized flow diagram of ground-water sampling steps (adapted from Barcelona et al., 1985).
134
-------
Table 9-7. Sources of Information on Various Aspects of Ground-Water sampling
Topic
References
Analyte identification
Well construction
Purging
Sample devices
Chemical changes
Comparisons
Packer samplers
Hydropunch
Discrete point
Sampling procedures
Decontamination
Metals
Volatiles
Oil-water mixtures
Field measurement
Barcelona (1983), Battista and Connelly (1989), Spruill (1988)
Alter et at. (1989), Cohen and Rabold (1988), Hackett (1988), Palmer at at. (1987), Pennine (1988), Perry and Hart
(1985), Sykes et at. (1986). See also Table 9-12
Barcelona and Helrich (1986), Barber and Davis (1987), Gibs and Imbrigiotta (1990), Herzog et al. (1988),
Oliveros et al. (1988), Palmer et al. (1987), Panko and 8arth (1988), Pennine (1988), Robbins (1989), Robin
and Gillham (1987), Smith et al. (1988), Unwin and Maltby (1988). See also, Table 9-11
Barcelona et al. (1985b), Barker and Dickhout (1988), Holm et al. (1988), Pannino (1988), Stolzenburg and
Nichols (1985), Schalla et al. (1988), Rose and Long (1988)
Barcelona et al. (1984), Barcelona et al. (1988), Pohlmann and Hess (1988), Nielsen and Yeates (1985)
Anderson (1979)
Cordry (1986), Edge and Cordry (1989)
McPherson and Pankow (1988)
Meade and Ellis (1985), Mickam et al. (1989)
Puts and Barcelona (1989)
Barker and Dickhout (1988), Schalla et al. (1988), Unwin and Maltby (1988)
Borst (1987)
Garner (1988), Garske and Schock (1986), Holm et al. (1987)
Table 9-8. Chemical Constituents of Interest in Ground- Water Monitoring
Type of Analyte
Geochemical
pH.Eh
Conductivity
Temperature
Dissolved oxygen
Alkalinity
Ca", Mg"
A/a', K'
Cr, SO/, PO;
Silicate
Where Done
L = Lab
F = Field
FF= Field
Filtered
F
F
F
F
F(FF)
L(FF)
L(FF)
F(FF)
L(FF)
Information Applications
Water
quality
X
X
X
X
X
X
Drinking
water
X
X
X
X
X
Contami-
nation
X
X
X
X
X
Possible
source
impacts
X
X
X
X
X
Geochemical
evaluation
of data
X
X
X
X
X
X
X
X
Water quality
Trace Metals
(Fe, Mn, Cr
Cd, Pb, Cu)
NO,, NH/
F
roc
TOX
TDS
Organic
compounds
L(FF)
L(FF)
L
L
L
L(FF)
L
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
Source: Modified from Barcelona et al. (1989)
135
-------
Table 9-9. Recommended Analytical Parameters for Detective Monitoring
Analytes
Type of Parameter
Well-purging
Contamination
indicators
Water quality '
Where Measured
F = Field, L = Lab
F
F
L
L
L
L
Required by regulation
pH, conductivity
pH, conductivity
TOC (total organic carbon)
TOX (total organic halogen)
Cl ; Fe, Mn, A/a •, SO/
Phenols
Suggested for Completeness
Temperature
Redox potential
Alkalinity (F) or acidity (F)
Ca", Mg", K>,
Drinking water
suitability "
As, Ba, Cd, Cr, F, Pb, NO,, Se, Ag
Endrin, lindane, methoxychlor,
toxaphene, 2,4-D, 2,4,5-TP (Silvex)
Radium, gross alpha/beta
Coliform bacteria
' All parameters required to be determined quarterly for the first year of network operations (RCRA Pan 265.92).
* These parameters are excluded from the annual reporting requirements of RCRA after the first year.
Source: Barcelona et al. (1985a)
9.3.3 Purging
Purging involves removing stagnant water from a moni-
toring well before taking a sample for analysis. Once monitor-
ing well locations have been selected, inadequate purging
procedures probably account for more sampling error than
any other step of the sampling process (see Table 7-2), There
is no universally correct purge volume. Monitoring wells
finished in materials of widely varying hydraulic conductivity
may require different purge volumes since chemical constitu-
ents are likely to migrate towards a pumped well at different
rates (Gibs and Imbriggiotta, 1990).
Recommended rules of thumb such as using 3 to 5
volumes (Fenn et al., 1977) should be treated only as a starting
point. Consistent estimation of purge volume requires know-
ing (1) well yield, determined from a slug or pumping test
and (2) the stagnant volumes of both the well casing and the
sand pack. Pumping rates for purging (i.e., generally 1 to 5
gpm) should be below the rate used for development (gener-
ally 5 to 20 gpm) to avoid well damage, which could induce
the migration of fines into the screened interval. The length of
time required to remove the stagnant water at the planned
pumping rate can readily be estimated from the well yield and
stagnant volume calculations. In most cases, it is important to
minimize the purge requirement to avoid dealing with large
volumes of contaminated water.
Monitoring pH, conductance, and temperature during
purge pumping can provide indications of background chem-
istry. After the stagnant water has been removed or isolated,
these indicators should continue to be monitored until they
reach a consistent end point (no upward or downward trend)
before sampling. Even after stagnant water has been removed,
some constituents may show increasing or decreasing trends.
Table 9-11 summarizes the results of observations in seven
studies where concentrations were measured as a function of
well volumes pumped, Increasing or decreasing concentration
trends usually will reach a constant level, although volatile
constituents may show considerable variance (see below).
The site- and constituent-specific nature of concentration
trends with purging is evident from the fact that bicarbonate,
nitrate, and specific conductance exhibited both increasing
and decreasing trends in different studies.
In studies by Smith et al. (1988), measurements of tri-
chloroethylene ranged from O to 250 mm as a function of
purge volumes from O to 25 volumes inside the well casing
and the volume inside the sand pack. After two to three well
volumes, trichloroethylene concentrations reached 100 to 125
ug/L. Five to ten well volumes averaged 150 to 175 ug/L, so
at least five well volumes was required to obtain samples near
the average. Concentrations dropped quickly after purging
stopped, and purging a day later yielded similar results. This
effect is probably the result of volatile losses from the stag-
nant water.
9.3.4 Well Construction and Sampling Devices
The Hydropunch® collects one-time ground-water samples
in unconsolidated material (see Figure 9-10). It is driven into
the soil and when the bottom of the probe is at least 5 ft below
the water table, the outer cylinder can be pulled back exposing
a perforated stainless steel sample entry barrel covered with
either a nylon or polyethylene filter material (see Figure 9-
lla). Hydrostatic pressure forces ground water that is rela-
tively free of turbidity into the sample compartment (see
Figure 9-llb). About 6 to 10 water samples of between 500
and 1,000 mL each often can be obtained in this manner if no
major problems occur. Geologic materials that can be augured
or sampled with a split spoon are suitable for sampling with
the Hydropunch®.
All decisions preceding monitoring well construction and
sample collection have to be quality-assured and documented.
Among the references listed in Table 9-1, Barcelona et al.
(1983) and Aller et al. (1989) focus primarily on monitoring
well design and construction. Screen slot size selection should
be justified, preferably by a quick sieve analysis in the field.
136
-------
Table 9-10. Chemical Constituents of Formation Waters Analyzed in Studies Related to Deep-Well Injection
Constituent
Depth (ft)
Temperature (C)
Specific gravity
pH
Eh
Conductance
TDS
Alkalinity
Pheno. alkalinity
Hardness
COD
Silica
Calcium
Magnesium
Sodium
Potassium
Bicarbonate
Sulfate
Chloride
Fluoride
Bromide
Iodide
Nitrite/nitrate
Ammonium (N)
Organic N
Orthophosphate
Hydrogen sulfide
DOC
Organic carbon
Acetate
Propionate
Buiyrate
Tola! Org. acids
Titrated org. alk.
Aluminum
Arsenic
Barium
Boron
Beryllium
Cadmium
Chromium
Cobalt
Copper
Iron (total)
Ferrous iron
Lead
Lithium
Manganese
Mercury
Molybdenum
Nickel
Selenium
Strontium
Zinc
Wilmington
NC
900
22.7
1.009
7.4
—
31,800
20,800
—
2,110
—
9
333
309
6,750
186
230
273
12,100
<1
—
—
<0.1
—
—
<0.1
tr.
<1
—
—
—
—
—
—
<7
<0.01
<1
—
—
<0.1
<0.1
<0.01
<0.1
2
—
<0.01
<1
<1
0.01
<0.01
<0.01
<0.01
19
<0.1
Pensacola
FL
1,430
35.2
—
7.4
-0.032
22,320
13,700
—
1,060
—
18
181
142
4,920
65
302
0
8,150
3
28
2
0
8
2
0
1
2
—
—
—
—
—
—
—
<0.1
—
5
—
—
—
—
-------
Table 9-11. Observed Trends in Measured Concentrations
with Well Volumes Pumped
References Indicating Trend in
Measured Concentration
Parameter Increasing
Arsenic 2
Alkalinity
Ammonium
Bicarbonate 5
Boron
Cadmium
Calcium
Carbonate
Chloride
Chromium
Copper
DOC 3
Hardness
Iron
Fluoride
Magnesium 2
Manganese
Nitrate 1,6,7
pH
Potassium
Selenium 2
Sodium
Specific 7
Conductance
Sulfate
TDS
Temperature 7
Zinc
Constant
1
1,3,5
2,3
2
1,2,3,5
1,3,5
5
1
2
3,5
7,2,3,5
2
f
1,3
2,3,5
1,2,3,5
1,3
1,2,5
1,3
Decreasing
I
3
1,3,5
3,5
7
2,5
1
2
5
4,6
1,5
1,3,4,6
5
f
2,5
References:
1 Chapin (1981)
2 Gibb et al. (1981)
3 Slawson et al. (1982)
4 Schmidt (1982)
5 Marsh and Lloyd (1980)
6 Nightingale and Bianchi (1980)
7 Keith et al. (1982)
Drilling or Penetrometer Rod
Upper Check Valve
' Sample Discharge Port
Adaptor to Drilling or
Penetrometer Rod
Sample Chamber
Lower Check Valve
*jh Slide Assembly
• - Sample Intake Tube
I
'£ T O Ring
Dr/Ve Cone
Figure 9-10. HydroPunch® schematic (from Edge and Cordry,
1989).
Source: Adapted from Rehm et al. (1985)
Common rigid well-casing materials that might be used in-
clude polyvinyl chloride, stainless steel, and polytetrafluoro-
ethylene. Table 4-3 summarizes the advantages and
disadvantages of these and other well casing and screen
materials. Figure 9-12 shows a sample decision tree for the
selection of rigid materials for casing.
Table 9-12 summarizes data on the leaching and sorption
characteristics of well casing materials. Stainless steel may be
the best overall metal easing and screening material, but it is
still susceptible to microbiological corrosion. In most in-
stances, casing and screen materials should last for at least 30
years without corrosion closing down the effective area of the
screen. Teflon® and polyvinyl chloride have structural prob-
lems for emplacement in deeper holes. All common easing
and tubing materials may be expected to sorb hydrophobic
organics to some extent. The impact of sorptive losses or
leaching contamination can be expected to be different with
aged materials than with the virgin material.
Figure 9-13 summarizes recommended sampling meth-
ods for various parameters for detective monitoring programs
and Figure 9-14 shows a decision tree for selection of sam-
pling mechanisms. With sampling devices, pressure changes
and the loss of volatiles are the main concern. Sampling
within 30 feet of the surface involves little pressure change
and most samplers may be expected to perform similarly for
volatile and gas-sensitive species. Sampling at depths in ex-
cess of 60 ft (two or more atmospheres) can be expected to
yield differences in sampling devices. Teflon®, polypropylene,
and polyethylene are the best tubing materials for sampling.
Polyvinyl chloride, Tygon®, and silicone rubber tubing should
138
-------
Penetrometer Rods
Hydropunch
Groundwater Sampler
v Water Table
Soil
Grounwater
Flow Path
Figure 9-11.
HydroPunch® sampling operation.
(a) HydroPunch® is pushed Into target ares with a
cone-penetrometer rig. (b) Once exposed, ground
water flows through the intake tube and into the
sample chamber (from Edge and Cordry, 1989).
be avoided, particularly if VOCs are involved, due to docu-
mented major losses of these species. Dedicated sampling
devices can greatly increase the cost efficiency of taking
samples.
A bladder pump is a cylinder with an internal bladder that
can be compressed and expanded under the influence of a gas.
The squeezing and release of pressure can be controlled with a
frequency that will give virtually pulseless flow. Bladder
pumps operate on air or nitrogen and air compressors are
available that are relatively easy to move around for supplying
them. Bladder pumps provide precise flow rates at given
operating pressures and frequencies of pressure/release. They
have worked reliably with continuous submersion in the same
well for extended periods. Any malfunction such as a leak in a
bladder pump is immediately apparent because it will stop
working. Repair in the field is also relatively easily accom-
plished. Bladder pumps are best adapted for purging small-
diameter monitoring wells (less than 4 in.) and their depth
range is limited to about 450 ft.
Bailers are commonly used sampling devices, but have a
number of disadvantages compared to bladder pumps. The
basic performance difficulties with bailers are that virtually all
individuals bail differently, and in-line determination of pH,
conductance, temperature, and dissolved oxygen are not pos-
sible. Also, sample transfer can be inconsistent, which creates
variability that shouldn't be in the sample data set. Another
major problem with bailers is the difficulty of determining
where a sample is actually retrieved. In this case, bailers may
malfunction without the operator knowing when the check
valve actually sealed. Bailers or grab samplers can minimize
volatilization losses, and are probably the best way to sample
NAPLs at the water table surface. Newer bailer designs allow
filtration in the field and transfer of volatile samples without
contact with the atmosphere, but to not address the problem of
inconsistency in bailing. Bailers should not be used for purg-
ing because all they do is homogenize the volume within the
well bore.
Electric submersibles can be useful for purging large-
diameter deep wells with high volume purging requirements,
particularly when flow rates can be controlled. They may not
be good for sampling unless the flow rate for sampling can be
controlled or diverted from the main pumping stream. In
general the accuracy is poor for gas-sensitive parameters, not
only volatile organics, but also oxygen and carbon dioxide.
Suction pumps, venturi mechanism pumps, and some
grab-driven mechanisms create turbulence that puts negative
pressure on the sample for volatiles. Flow is generally diffi-
cult to control, particularly to obtain preferable low flows
(i.e., 100 mL to 2 L/min) for sampling.
Sampling devices and sample handling should be ex-
ecuted so as to minimize temperature and pressure changes.
Reproducible flow rates and freedom from operator-induced
errors tend to yield the most precise results. Figure 9-15
contains a matrix rating the suitability of different chemical
devices for different chemical constituents. Figure 9-16 rates
suitability of 12 devices (described in Table 9-13) for use with
12 types of ground-water parameters. If VOCs are sampled
139
-------
Rigid Material Recommendation
Purpose of Program
••••••••••I
L
Detective
Assessment
Subsurface
Contamination
Conditions
Unknown
PTFE
SS
PVC
High Organic
High Inorganic or Inorganic
Suspected Suspected
PTFE
SS
PVC
Importance
of Trace
Level
Organics
High Organic
No Inorganic
Known
PTFE
PVC
SS
High Organic
and/or
Inorganic
Known
PTFE
SS
Could PPB Level Organics be
Important?
Yes
PTFE=polytetrafluoroethylene
SS = stainless steel (316 or 304)
PVC = polyvinyl chloride
PTFE
SS
PTFE
SS
PVC
Figure 9-12. Decision tree for recommended well-casing/screen materials. Adapted by Barcelona and Gibb (1888) from Barcelona
et al. (1985).
140
-------
Table 9-12.
Effects of Well Casing Material! on Trace
Concentration in Well Water
Parameter
Arsenic
Cadmium
Chromium
Copper
Dissolved Organic
Carbon
iron
Manganese
Total Organic Carbon
Zinc
Lithium
Mercury
Molybdenum
Selenium
Leaches From
ABS*
Steel, Galvanized 2
Steel
Steel
ABS, PVC**
Steel, Galvanized '
Steel, Galvanized1
ABS, PVC
Steel, Galvanized '
Adsorbed By
ABS
ABS
ABS
ABS
ABS
ABS
Source: Adapted from Houghton and Berger (1984)
' Acrylonjtrile-butadiene-styrene copolvmer
" Polyvinyl chloride
1 Suggested by data from Gibb et al. (1981)
2 Barcelona et al. (1983)
effectively, the results with most other constituents may be
expected to be reproducible and accurate.
Studies of sampling errors associated with the sampling
mechanisms alone have found that bladder pumps and bailers
come out with sampling error consistently less than the
analytical error. However, most comparisons of bladder pumps
and bailers have been conducted at shallow lifts. At depths up
to 200 or 300 m, bladder pumps are probably superior.
Vacuum devices, peristaltic and suction pumps, on the other
hand, yield a sampling error on about the same order as the
level of sorptive losses or handling errors. Where sensitive
constituents are involved, bladder pumps and bailers are the
most frequently used devices. A bladder pump is probably
the best overall sampling device and will probably provide
50 to 100 percent better recovery and far better precision for
volatiles than a bailer.
Once samples are collected, procedures for the handling
and preservation of samples should be carefully followed to
minimize errors from this stage of the sampling process.
Table 9-14 summarizes recommended sample handling and
preservation procedures for a comprehensive detective moni-
toring program.
141
-------
Hydrogeologic Conditions (yield capability)
Parameters
(type)
Well-purging
(pH,Eh.T,n-<)
Contamination
Indicators
(pH,n<)
(TOC, TOX)
Water Quality
Dissolved Gases
(Of, CH4, COJ
Alky/Acdy
(Fe,Mn,P04;CI;
A/a*, SO/, Ca",
Mg", K; NO,,
Silicate
(Ammonium,
Phenols)
Drinking Water
Suitability
(As, Ba, Ctf, Cr, Pb,
Hg, Se.Ag, A/O,, F)
(Remaining
Parameters)
Mechsnism
(material)'
Pump
(T. S. P, 0)
Flow rates:
0,1-1, OUmin
Grab
(T, S, G, P, O)
Pump
Flow rates:
0.1-1. OUmin
Grab
(T, S, G, P, O)
Pump
(T, S preferred;
O,P only where
supporting data
exist)
Grab
(T, S, G preferred;
O, P only where
supporting data
exist)
Pump
(T, S, P, O)
Grab
(T, S, G, P, O)
Pump
(T, S, P, O)
Grab
(T, S, G, P, O)
Pump
(T, S preferred; O, P
only where support-
ing data exist)
Grab
(T, S, G preferred; O, P
only where support-
ing data exist)
Pump
(T, S, P. O)
Grab
(T, S, G, P, O)
Pump
(T, S, P, O)
Grab
(T, S, G, P, O)
(both with precautions
ifradiologic hazards
exist)
>100 mUmin yield
Flowing samples
Positive displacement
bladder pump (air, N2)
Positive displacement
bladder pump.
(air, Nt)
(Mechanism as above
operated ai flow rates noi
to exceed 100 mUmin)
Vials or bottles filled gently
from bottom up and al-
lowed to overflow-tTeflon
capped w/o headspace
(Mechanism as above
operated at flow rates not
to exceed 100 mUmin)
Glass containers filled
gently from bottom up and
allowed to overflow-tTeflon
capped w/o headspace
Positive displacement
bladder pump, (air, N2)
(Mechanisms as above
operated at flow rates not
to exceed 1000 mUmin)
Glass containers filled from
bottom up
Positive displacement
bladder pump, (air, N2)
Positive displacement
bladder pump, (air NJ
Flow rates should not ex-
ceed 1000 mUmin
<100 mUmin yield
Discrete samples
Dual check valve bailers
"thief" samplers
bladder pump (air, N2)
Dual check valve bailers
"thief" samplers
(Volatile fractions of TOC
and TOX may be lost de-
pending on conditions and
operator skill)
40-mL vials (500-mL Tef-
lon-sealed glass bottles for
TOX) filled from bottom up
and allowed to overflow or
gently poured down the
side of the vial. Teflon
capped w/o headspace
(Not recommended)
Fe values sensitive to most
grab mechanisms
Large volumes required
may have to be sequen-
tially collected
(Volatile species may be
lost depending on conditions
Glass containers filled from
bottom up
Dual check valve bailers
"thief" samplers
(Volatile compounds may
be lost depending on conditions)
"Materials in order of preference include: Teflon® (T); stainless steel (S): PVC, polypropylene, polyethylene (P); borosilicate glass (G); other
materials: silicone, polycarbonate, mild steal, etc. (0)
Figure 9-13. Recommended sample collection methods for detective monitoring programs (from Barcelona et al., 1985).
142
-------
Sampling Mechanism Recommendation
Lift Requirements
Flow Rate
Variability of Mechanism
(Purging and Sampling)
Yes
_L
<140m(450')
>140m
Limited
Grab or
Gas Lift
Pumps
Limited
Grab or Gas Lift
No Gas Contact
P-D Bladder
P-D Mechanical
Are
Parameters of
Interest
Volatile,
or pH Sensitive?
No Gas Contact:
P-D Bladder
P-D Mechanical
Gas-Drive
Centrifugal
Peristaltic
Suction
No
Yes
Yes
Bailer
Thief
Thief
No No
- Incomplete -
Yes
Figure 9-14. Decision tree for recommended purge and sampling mechanism. Adapted by Barcelona and Gibb (1988) from
Barcelona et al. (1985).
143
-------
Type of
Constituent
Volatile
Organic
Compounds
Organometallics
Dissolved Gasses
Well-Purging
Parameters
Trace Inorganic
Metal Species
Reduced
Species
Major Cations
& Anions
Example of
Constituent
Chloroform
TOX
CH3Hg
Oz, COz
pH, Q'1
Eh
Fe, Cu
Na+, K\ Ca++
Mg"
CI',SO4 =
Positive
Displacement
Bladder Pumps
| Increasing Sample Sensitivity »•
Superior
performance
for most
applications
Superior
performance
for most
applications
Superior
performance
for most
applications
Superior
performance
for most
applications
Thief, in Situ
or Check Valve
Bailers
Mechanical
Positive
Displacement
Pumps
Gas-drive
Devices
Suction
Mechanisms
Increasing Reliability of Sampling Mechanisms
May be adequate
if well purging
is assured
May be adequate
if well purging
is assured
May be adequate
if well purging
is assured
Adequate
May be adequate
if well purging
is assured
May be adequate
if design and
operation are
controlled
May be adequate
if design and
operation are
controlled
Adequate
Adequate
Not
recommended
Not
recommended
Maybe
adequate
Adequate
Not
recommended
Not
recommended
Maybe
adequate if
materials are
appropriate
Adequate
Figure 9-15. Matrix of sensitive chemical constituents and various sampling mechanisms (from Barcelona et al., 1985).
144
-------
Portable Sampling Devices *
1
6
Positive Displacement
(submersible)
Jii*
Gas Contact
In Situ
Sampling
Devices*
Device
Open bailer
Point-source
bailer
Syringe
sampler
Gear-drive
Bladder
pump
Helical rotor
Piston pump
(gas-drive)
Centrifugal
Peristaltic
Gas-lift
Gas-drive
Pneumatic
Approximate
Maximum
Sample Depth
No Limit
No Limit
No Limit
200ft
400 ft
160 ft
500ft
Variable
26ft
Variable
150ft
No Limit
Minumum
Well
Diameter
1/2 in
1/2 in
1 1/2 in
2 in
1 1/2in
2 in
1 1/2'm
3 in
1/2 in
1 in
1 in
Not
Applicable
Sample t
Delivery Rate
or Volume
Variable
Variable
0.01-0.2 gal
0-0.5 gpm
0-2 gpm
0-1. 2 gpm
0-0.5 gpm
Variable
0.01 -0.3 gpm
Variable
0.2 gpm
0.01-0.13 gal
Ground Water Parameters
Inorganic
EC
*
*
*
*
•
•
*
*
*
*
pH
•
•
•
•
•
Redox
•
*
•
•
•
Major
Ions
•
•
•
•
•
•
•
•
•
•
Trace
Metals
•
•
•
•
•
•
•
Nitrate,
Fluoride
•
•
•
•
•
•
•
•
•
•
Dis-
solved
Gasses
•
•
Organic
Non-
volatile
*
«
*
•
•
•
•
•
*
Volatile
•
•
*
•
TOC
•
•
•
TOX
9
•
*
Radioactive
Radium
•
•
•
•
•
•
•
•
•
•
Gross
Atfa&
Beta
•
•
•
•
•
•
Biol.
Coli-
form
Bacteria
•
•
•
•
•
•
* Sampling devices on this chart are divided into two categones: (1) portable devices for sampling existing monitoring
wells; and (2) in situ monitoring devices (often multilevel) that are permanently installed. Sampling device construction
materials (including tubing, haul lines, etc.) should be evaluated for suitability in analyzing specific ground water
parameters. It is assumed on this chart that existing monitoring wells are properly installed and constructed of materials
suitable for detection of the parameters of interest. See references for additional information.
t Sample delvivery rates and volumes are average ranges based on typical field conditions. Actual delivery rates area
function of diameter of monitoring installation, size and capacity of sampling device, hydrogeologic conditions, and depth
to sampling point. For all devices, delivery rate should be carefully controlled to prevent aeration or degassing of the
w sample.
Z Indicates device is generally suitable for application (assuming device is cleaned and operated properly and is
constructed of suitable materials).
Figure 9-16. Generalized ground-water sampling device matrix (from Pohlmann and Hess, 1988).
-------
Table 9-13. Description of Ground-Water Sampling Devices and Construction Materials Commonly Used in Ground- Water
Monitoring (see also Figure 8-16)
Sample Device Description
Open top. Bottom sealed or fitted with foot valve. Available in wide range of rigid materials.
Check valve at both top and bottom. Valves are opened by cable operated from ground surface.
Available in wide range of rigid materials.
Sample container is pressurized or evacuated and lowered into sampling installation. Opening the
container and/or releasing the pressure allows sample to enter the device. Materials may include
stainless steel 316, Teflon®, polyethylene, glass.
Electric motor rotates a set of Teflon gears, which drives the sample up the discharge line. Constructed
of stainless steel 304, Teflon, and Viton®.
Flexible bladder within device has check valves at each end. Gas from ground surface is cycled between
bladder and sampler wall, forcing sample to enter bladder and then be driven up the discharge line. Gas
does not contact sample. Materials may include stainless steel 316, Teflon, Viton, polyvinyl chloride (PVC),
silicone, Neoprene®, polycarbonate, Delrin®.
Water sample is forced up discharge line by electrically driven rotor-stator assembly. Materials may include
stainless steel 304, ethylene propylene rubber (EPDM), Teflon, Viton, polypropylene.
Piston is driven up and down by gas pressure controlled from the surface. Gas does not contact sample.
Materials may include stainless steel 304, Teflon, Delrin, polypropylene, Viton, acrylic, polyethylene.
Electrically driven rotating impeller accelerates water within the pump body, building up pressure and forcing
the sample up discharge line. Commonly constructed of stainless steel, rubber, and brass.
Self priming vacuum pump is operated at ground surface and is attached to tubing, which is lowered to the
desired sampling depth. Sample contacts vacuum. Materials may include Tygon®, silicone, Viton, Neoprene,
rubber, Teflon.
Gas emitted from gas line at desired depth forces sample to surface through sampling installation.
Another method utilizes gas to reduce effective specific gravity of water, causing it to rise.
Wide variety of materials available for tubing.
Positive gas pressure applied to water within device's sample chamber forces sample to surface. Materials
may include polyethylene, brass, nylon, aluminum oxide, PVC, polypropylene.
in situ device generally utilizes the same operating principles as syringe samplers: a pressurized or evacuated
sample container is lowered to the sampling port and opened, allowing the sample to enter. Materials may
include PVC, stainless steel, polypropylene, Teflon.
Source: Pohlmann and Hess (1988)
Open bailer
Point-source bailer
Syringe sampler
Gear-drive pump
Bladder pump
Helical-rotor pump
Gas-driven pump
Centrifugal pump
Peristaltic pump
Gas-lift devices
Gas-drive devices
Pneumatic
146
-------
Table 9-14. Recommended Sample Handling and Preservation Procedures for a Detective-Monitoring Program
Volume
Parameters Required (mL)
(Type) 1 Sample'
Well purging
pH (grab)
ii ' (grab)
T(grab)
Eh (grab)
Contamination
indicators
pH, n-' (grab)
TOC
TOX
Water quality
Dissolved gases
(0 CH4,COJ
Alkalinity/acidity
(Fe, Mn, A/a*,
K', Ca",
Mg»)
(P0t, Cr,
Silicate)
NO3
S04
OH4-
Phenols
Same as
suitability
As, Ba, Cd, Cr,
Pb.Hg.Se.Ag
F
50
100
1,000
1,000
As above
40
500
10 mL minimum
100
Filtered under
pressure with
appropriate
media
All filtered
1,000 mL '
@50
100
50
400
500
Same as
above for
water
quality
cations
(Fe, Mn,
etc.) '
Same as
chloride above
Containers
(Material)
T,S,P,G
T.S.P.G
T,S,P,G
T,S,P,G
As above
G,T
G,T
G,S
T.G.P
T.P
(T.P.G
glass only)
T.P.G
T.P.G
T.P.G
T.G
Same as
above
Same as
above
Preservation
Method
None: field del.
None: field del
None: field det.
None: field det.
As above
Dark, 4 °C
Dark, 4 °C
Dark, 4 °C
4 "C/None
Field acidified
to pH <2 with
HNG3
4°C
4°C
4°C
4 °C/HSO4 to
pH<2
4 °C,'HJPQ4 to
pH <4 Drinking Water
6 months
Same as above
Maximum
Holding
Period
<1 hr"
<1 hr"
None
None
As above
24 hi"
5 days
<24hr
<6hi*
<24hr
6 months'
24 hr/
7 days':
7 days
24 hr"
7 days'
24 hr/
7 days
24 hr
7 days
Remaining organic
parameters
As for TOX/TOC, except where analytical method calls for
acidification of sample
24 hr
"It is assumed that at each site, for each sampling date, replicates, a field blank, and standards must be taken at equal volume to those of the
samples.
' Temperature correction must be made for reliable reporting. Variations greater than ±10% may result from a longer holding period.
cln the event that HNO,cannot be used because of shipping restrictions, the sample should be refrigerated to 4°C, shipped immediately, and
acidified on receipt at the laboratory. Container should be rinsed with 1:1 HNO,and included with sample.
"28-day holding time if samples are preserved (acidified).
'Longer holding times in EPA (1986b).
'Filtration is not recommended for samples intended to indicate the mobile substance lead. See Puts and Barcelona (1989) for more specific
recommendations for filtration procedures involving samples for dissolved species.
Note: T= Teflon; S = stainless steal; P = PVC, polypropylene, polyethylene;
G = borosilicate glass.
Source: Adapted from Scalfet al. (1981) and U.S. EPA (1986b)
147
-------
9.4 References
Aller, L., T.W. Bennett G. Hackett, Rebecca J. Petty, J.H.
Lehr, H. Sedoris, D.M. Nielsen. 1989. Handbook of
Suggested Practices for the Design and Installation of
Ground-Water Monitoring Wells. EPA/600/4-89/034
(NTIS PB90-159807). Also published in NWWA/EPA
series, National Water Well Association, Dublin, OH.
Andersen, L.J. 1983. Sampling Techniques of Groundwater
from Water Wells. In: Proc. UNESCO Symp. Methods
and Instrumentation for the Investigation of Groundwater
Systems, Committee for Hydrological Research, CHO-
TNO, The Hague, The Netherlands, pp. 521-527.
Armstrong, J.M., W. Korreck, L.E. Leach, R.M. Powell, S.V.
Vandegrift, and J.T. Wilson. 1988. Bioremediation of a
Fuel Spill: Evaluation of Techniques for Preliminary Site
Characterization. In: Proc. 5th NWWA/API Conf. Petro-
leum Hydrocarbons and Organic Chemicals in Ground
Water-Prevention, Detection and Restoration, National
Water Well Association, Dublin, OH, pp. 931-948.
Barber, C. and G.B. Davis. 1987. Representative Sampling of
Ground Water from Short-Screened Boreholes. Ground
Water 25(5):581-587.
Barcelona, M.J. 1983. Chemical Problems in Ground-Water
Monitoring Programs. In: Proc. 3rd Nat. Symp. on Aqui-
fer Restoration and Ground Water Monitoring, National
Water Well Association, Dublin, OH, pp. 263-271.
Barcelona, M.J. and J.P. Gibb. 1988. Development of Effec-
tive Ground-Water Sampling Protocols. In: Ground-Wa-
ter Contamination: Field Methods, A.G. Collins and A.L
Johnson (eds.), ASTM STP 963, American Society for
Testing and Materials, Philadelphia, PA, pp. 17-26.
Barcelona, M.J. and J.A. Helfrich. 1986. Well Construction
and Purging Effects on Ground-Water Samples. Environ.
Sci. Technol. 20:1179-1184.
Barcelona, M.J., J.P. Gibb, and R.A. Miller. 1983. A Guide to
the Selection of Materials for Monitoring Well Construc-
tion and Ground-Water Sampling. ISWS Contract Report
327. Illinois State Water Survey, Champaign, IL.
Barcelona, M.J., J.A. Helfrich, E.E. Garske, and J.P. Gibb.
1984. A Laboratory Evaluation of Ground Water Sam-
pling Mechanisms. Ground Water Monitoring Review
4(2):32-41.
Barcelona, M.J., J.P. Gibb, J.A. Helfrich, and E.E. Garske,
1985a. Practical Guide for Ground-Water Sampling. EPA/
600/2-85/104 (NTIS PB86-137304). Also published as
ISWS Contract Report 374, Illinois State Water Survev,
Champaign, IL.
Barcelona, MI, J.A. Helfrich, and E.E. Garske. 1985b. Sam-
pling Tubing Effects on Groundwater Samples. Anal.
Chem. 57:460-464.
Barcelona, MI, J.A. Helfrich, and E.E. Garske. 1988. Verifi-
cation of Sampling Methods and Selection of Materials
for Ground-Water Contamination Studies. In Ground-
Water Contamination: Field Methods, A.G. Collins and
A.I. Johnson (eds.), ASTM STP 963, American Society
for Testing and Materials, Philadelphia, PA, pp. 221-231.
Barcelona, M.J., D.P. Lettenmaier, and M.R Schock. 1989.
Network Design Factors for Assessing Temporal Vari-
ability in Ground-Water Quality. Environmental Moni-
toring and Assessment 12:149-179.
Barker, J.F. and R. Dickhout. 1988. An Evaluation of Some
Systems for Sampling Gas-Charged Ground Water for
Volatile Organic Analysis. Ground Water Monitoring
Review 8(4): 112-119.
Earth, D.S. and T.H. Starks. 1985. Sediment Sampling Qual-
ity Assurance User's Guide. EPA/600/4-85&8 (NTIS
PB85-233542).
Earth, D.S., B.J. Mason, and T.H. Starks. 1989. Soil Sampling
Quality Assurance User's Guide, 2nd ed. EPA/600/8-89/
046 (NTIS PB89-189864), 225 pp.
Battista, J.R. and J.P. Connelly. 1989. VOC Contamination at
Selected Wisconsin Landfills-Sampling Results and
Policy Implications. WDNR PUBL-W-094 89. Wiscon-
sin Department of Natural Resources, Madison, WI.
Bell, H.F. and H.P. DeLong. 1988. Data Characteristics Ground
Water Monitoring's Catch 22. ACS Abstracts 28(2):20-
24.
Berg, E.L. 982. Handbook for Sampling and Sample Preser-
vation of Water and Wastewater. EPA/600/4-82-029
(NTIS PB83-124503).
Borst, M. 1987. Sampling Oil-Water Mixtures at OHMSEIT.
EPA/600/2-87/073 (NTIS PB88-102892).
Brown, K.W. 1987. Efficiency of Soil Core and Soil-Pore
Water Sampling Systems. EPA/00/2-86/083 (NTIS PB87-
106100).
Bumb, A.C., C.R. McKee, R.B. Evans, and LA. Eccles. 1988.
Design of Lysimeter Leak Detector Networks for Surface
Impoundments and Landfills. Ground Water Monitoring
Review 8(2): 102-114.
Cameron, R.E., et al. 1966. Sampling and Handling of Desert
Soils. Technical Report No. 32-908. Jet Propulsion Labo-
ratory, California Institute of Technology, Pasadena, CA,
37pp.
Chapin, R.I. 1981. Short-Term Variations, Sampling Tech-
niques and Accuracy of Analysis of the Concentrations of
Nitrate in Produced Municipal Ground Waters—North
Texas. M.A. Thesis, University of Texas, Austin.
148
-------
Close, M.E. 1989. Effect of Serial Correction on Ground
Water Quality Sampling Frequency. Water Resources
Bulletin 25(3):507-515.
Cochran, W.G. 1977. Sampling Techniques, 3rd ed. John
Wiley & Sons, New York.
Cohen, R.M. and R.R. Rabid. 1988. Simulation of Sampling
and Hydraulic Tests to Assess a Hybrid Monitoring Well
Design. Ground Water Monitoring Review 8(l):51-59.
Collins, A.G. and A.I. Johnson (eds.). 1988. Ground-Water
Contamination: Field Methods. ASTM STP 963. Ameri-
can Society for Testing and Materials, Philadelphia, PA.
Cordry, K. 1986. Ground Water Sampling Without Wells. In:
Proc. 6th Nat. Symp. on Aquifer Restoration and Ground
Water Monitoring, National Water Well Association,
Dublin, OH, pp. 266-271.
Deming, W.E. 1950. Some Theory of Sampling. John Wiley
& Sons, New York, (available as a Dover reprint, 1966).
Devitt, D. A., R.B. Evans, W.A. Jury, T.H. Starks, and B.
Eklund. 1987. Soil Gas Sensing for Detection and Map-
ping of Volatile Organics. EPA/600/8-87/036 (NTIS
PB87-228516).
Di, H.J., B.B. Trangmar, and R.A. Kemp. 1989. Use of
Geostatistics in Designing Sampling Strategies for Soil
Survey. Soil Sci. Soc. Am. 153:1163-1167.
Diem, D.A., B.E. Ross and H.B. Kerfoot. 1988. Field Evalua-
tion of a Soil-Gas Analysis Method for Detection of
Subsurface Diesel Fuel Contamination. In: Proc. 2nd Nat.
Outdoor Action Conf. on Aquifer Restoration, Ground
Water Monitoring and Geophysical Methods, National
Water Well Association, Dublin, OH, pp. 1015-1031.
Dunlap, W.J., J.F. McNabb, M.R. Scalf, and R.L. Cosby.
1977. Sampling for Organic Chemicals and Microorgan-
isms in the Subsurface. EPA/600/2-77/176 (NTIS PB272
679).
Edge, R.W. and K. Cordry. 1989. The Hydropunch: An In
Situ Sampling Tool for Collecting Ground Water from
Unconsolidated Sediments Ground Water Monitoring
Review 9(3):177-183.
Everett, L.G., K.D. Schmidt, R.M. Tinlin, and O.K. Todd.
1976. Monitoring Groundwater Quality: Methods and
costs. EPA/600/4-76/023 (NTIS PB257133).
Everett, L.G. and L.G. McMillion. 1985. Operational Ranges
for Suction Lysimeters. Ground Water Monitoring Re-
view 5(3):5 1-60.
Everett, L.G., L.G. Wilson, and L.G. McMillion. 1982. Va-
dose Zone Monitoring Concepts for Hazardous Waste
Sites. Ground Water 20(3):312-324,
Everett, L.G., L.G. Wilson, and E.W. Hoylman. 1983. Vadose
Zone Monitoring for Hazardous Waste Sites. EPA/600/
X-83/064 (NTIS PB84-212752). Also published in 1984
by Noyes Data Corporation, Park Ridge, NJ.
Everett, N. G., E.W. Hoylman, L.G. Wilson, and L.G.
McMillion. 1984. Constraints and Categories of Vadose
Zone Monitoring Devices. Ground Water Monitoring
Review 4(1):26-31.
Fenn, D., E. Cocozza, J. Isbister, O. Braids, B. Yare, and P.
Roux. 1977. Procedures Manual for Ground Water Moni-
toring at Solid Waste Disposal Facihties.EPA/530/SW611
(NTIS PB84-174820).
Flatman, G.T. 1986. Design of Soil Sampling Programs:
Statistical Considerations. In: Quality Control in Reme-
dial Site Investigation: Hazardous and Industrial Solid
Waste Testing, 5th volume, C.L. Perket (ed.), ASTM
STP 925, American Society for Testing and Materials,
Philadelphia, PA, pp. 43-56.
Ford, P.J., P.J. Turina, and D.E. Seely. 1984. Characterization
of Hazardous Waste Sites—A Methods Manual: Vol. II.
Available Sampling Methods, 2nd ed. EPA/600/4-84-076
(NTIS PB85-521596).
Garner, S. 1988. Making the Most of Field-Measurable Ground
Water Quality Parameters. Ground Water Monitoring
Review 8(3):60-66.
Garske, E.E. and M.R. Schock. 1986. An Inexpensive Flow-
Through Ceil and Measurement System for Monitoring
Selected Chemical Parameters in Ground Water. Ground
Water Monitoring Review 6(3):78-84.
GeoTrans, 1989. Groundwater Monitoring Manual for the
Electric Utility Industry. Edison Electric Institute, Wash-
ington, DC.
Gibb, J.P, R.M. Schuller, and R.A. Griffh. 1981. Procedures
for the Collection of Representative Water Quality Data
from Monitoring Wells. ISWS/1GS Cooperative Ground
Water Report 7. Illinois State Water Survey, Champaign,
IL.
Gibs, J. and T.E. Imbrigiotta. 1990. Well-Purging Criteria for
Sampling Purgeable Organic Compounds. Ground Water
28(l):68-78.
Gilbert, R.O. 1987. Statistical Methods for Environmental
Pollution Monitoring. Van Nostrand Reinhold, New York.
Gillham, R.W., M.J.L. Robin, J.F. Barker and J.A. Cherry.
1983. Groundwater Monitoring and Sample Bias. API
Publication 4367. American Petroleum Institute, Wash-
ington, DC.
Glaccum, R., M. Noel, R. Evans, and L. McMillion. 1983.
Correlation of Geophysical and Organic Vapor Analyzer
Data Over a Conductive Plume Containing Volatile Or-
ganics. In Proc. 3rd Nat. Symp. on Aquifer Restoration
149
-------
and Ground Water Monitoring, National Water Well
Association, Dublin, OH, pp. 421-427.
Goodwin, C.D. and R.M. Burger. 1989. Inexpensive but Ef-
fective Soil Gas Screening—A Case Study. In: Proc. 3rd
Nat. Outdoor Action Conf. on Aquifer Restoration, Ground
Water Monitoring and Geophysical Methods, National
Water Well Association, Dublin, OH, pp. 263-271.
Goolsby, D. A. 1972. Geochemical Effects and Movement of
Injected Industrial Waste in a Limestone Aquifer. In:
Symposium on Underground Waste Management and
Environmental Implications (Houston, TX), T.D. Cook
(ed.), Am. Ass. Petr. Geol. Mere. 18, pp. 355-368.
Hackett G. 1988. Drilling and Constructing Monitoring Wells
with Hollow-Stem Augers: Part 2. Monitoring Well In-
stallation. Ground Water Monitoring Review 8(1)60-68.
Hansen, M.H. W.N. Hurwitz, and W.G. Madow. 1953. Sample
Survey Methods and Theory, Vols. I and H. John Wiley
& Sons, New York.
Hendricks, W.A. 1956. The Mathematical Theory of Sam-
pling. Scarecrow Press, New Brunswick, NJ.
Henzog, B.L., S.-F. Chou, J.R. Valkenberg, and R.A. Gnffm.
1988. Changes in Volatile Organic Chemical Concentra-
tions After Purging Slowly Recovering Wells. Ground
Water Monitoring Review 8(4): 93-99.
Hoffman, S.J. 1986. Soil Sampling. In: Exploration Geo-
chemistry, Design and Interpretation of Soil Surveys.
Reviews in Economic Geology 3:39-71.
Holden, P.W. 1984. Primer on Well Water Sampling for
Volatile Organic Compounds. Water Resources Research
Center, University of Arizona, Tucson, AZ.
Holm, T.R., G.K. George, and M.J. Barcelona. 1987. Fluoro-
metric Determination of Hydrogen Peroxide in Ground-
water. Anal. Chem. 59:582-586.
Holm, T.R., G.K. George and M.J. Barcelona. 1988. Oxygen
Transfer through Flexible Tubing and Its Effects on
Ground Water Sampling Results. Ground Water Moni-
toring Review 8(3): 83-89.
Houghton, R.L. and M.E. Berger. 1984. Effects of Well-
Casing Composition and Sampling Method on Apparent
Quality of Ground Water. In: Proc. 4th Nat. Symp. on
Aquifer Restoration and Ground Water Monitoring, Na-
tional Water Well Association, Dublin, OH, pp. 203-213.
Howsam, P. (ed.). 1990. Proceedings of International Ground-
water Engineering Conference on Water Wells: Monitor-
ing, Maintenance, and Rehabilitation. Chapman and Hall,
London, 422 pp.
Hsueh, Y-W. and R. Rajagopal. 1988. Modeling Ground
Water Quality Sampling Decisions. Ground Water Moni-
toring Review 8(4): 121-133.
Hughes, J.P. and D.P. Lettenmaier. 1981. Data Requirements
for Kriging: Estimation and Network Design. Water Re-
sources Research 17:1641-1650.
Jessen, R.J. 1978. Statistical Survey Techniques. John Wiley
& Sons, New York.
Johnson, T.M. and K. Cartwright. 1980. Monitoring of
Leachate Migration in the Unsaturated Zone in the Vicin-
ity of Sanitary Landfills. ISGS Circular 514. Illinois State
Geological Survey, Champaign, IL.
Kaufman, M. I., D.A. Goolsby, and G.L. Faulkner. 1973.
Injection of Acidic Industrial Waste into a Saline Carbon-
ate Aquifer Geochemical Aspects. In: Symposium on
Underground Waste Management and Artificial Recharge,
J. Braunstein (ed.), Int. Ass. of Hydrological Sciences
Pub. No. 110, pp. 526-551.
Keith, S.J., L.G. Wilson, H.R. Fitch, and D.M. Esposito.
1982. Sources of Spatial-Temporal Variability in Ground-
Water Quality Data and Methods of Control: Case Study
of the Cortaro Monitoring Program, Arizona. In: Proc.
2nd Nat. Symp. on Aquifer Restoration and Ground
Water Monitoring, National Water Well Association,
Dublin, OH, pp. 217-227.
Kerfoot, H.B. and J. Soderberg. 1988. Three-Dimensional
Characterization of a Vadose Zone Plume in Irregularly
Interbedded Silt and Sand Deposits. In: Proc. 2nd Nat.
Outdoor Action Conf. on Aquifer Restoration, Ground
Water Monitoring and Geophysical Methods, National
Water Well Association, Dublin, OH, pp. 1071-1087.
Kerfoot, H.B. and L.J. Barrows. 1987. Soil Gas Measurement
for Detection of Subsurface Organic Contamination. EPA/
600/2-87/027 (NTIS PB87-174884).
Kirschner, Jr., F.E. and G.L. Bloomsburg. 1988. Vadose Zone
Monitoring: An Early Warning System. Ground Water
Monitoring Review 8(2): 49-50.
Kish, L. 1965. Survey Sampling. John Wiley & Sons, New
York.
Kreitler, C.W., M.S. Akhter, and A.C.A. Donnelly. 1988,
Hydrologic-Hydrochemical Characterization of Texas
Gulf Coast Formations Used for Deep-Well Injection of
Chemical Wastes. University of Texas at Austin, Bureau
of Economic Geology.
Leenheer, J.A. and R.L. Malcolm. 1973. Case History of
Subsurface Waste Injection of an Industrial Organic Waste.
In: Symposium on Underground Waste Management and
Artificial Recharge, J. Braunsteirt (ed.), Int, Ass. of Hy-
drological Sciences Pub. No 110, pp. 565-579.
Leenheer, J. A., R.L. Malcolm, and W.R. White. 1976, Physi-
cal, Chemical and Biological Aspects of Subsurface Or-
ganic Waste Injection near Wilmington, North Carolina.
U.S, Geological Survey Professional Paper 987.
150
-------
Litaor, M.I. 1988. Review of Soil Solution Samplers. Water
Resources Research 24(5):727-733.
Loaiciga, H.A. 1989. An Optimization Approach for Ground-
water Quality Monitoring Network Design. Water Re-
sources Research 25(8): 1771-1782.
Loehr, R.C., J.H. Martin, Jr., E.F. Neuhauser. 1986. Spatial
Variation of Characteristics in the Zone of Incorporation
at an Industrial Waste Land Treatment Site. In: Hazard-
ous and Industrial Solid Waste Testing: Fourth Sympsium,
ASTM STP 886, J.K. Petros, Jr., W.J. Lacy, and R.A.
Conway (eds.), American Society for Testing and Materi-
als, Philadelphia, PA, pp. 285-297.
Mackay, D. and W.Y. Shiu. 1981. A Critical Review of
Henry's Law Constants for Chemicals of Environmental
Interest. J. Phy. Chem. Ref Data 10(4):1 175-1199.
MacPhemon. Jr., J.R. and J.F. Pankow. 1988. A Discrete Point
Sampler for Ground Water Monitoring Wells. Ground
Water Monitoring Review 8(3): 160-164.
Malcolm, R.L. and J.A. Leenheer. 1973. The Usefulness of
Organic Carbon Parameters in Water Quality Investiga-
tions. In: Proc. of the Inst. of Env. Sciences 1973 Annual
Meeting (Anaheim, CA), pp. 336-340. Available from
J.A. Leenheer, USGS MS 408, Box 25046, Federal Cen-
ter, Denver CO, 80225.
Marrin, D.L. 1987. Soil Gas Sampling Strategies: Deep vs.
Shallow Aquifers. In: Proc. 1st Nat. Outdoor Action
Conf. on Aquifer Restoration, Ground Water Monitoring
and Geophysical Methods, National Water Well Associa-
tion, Dublin, OH, pp. 437-454.
Marrin, D.L. and W.B. Kerfoot. 1988. Soil Gas Surveying
Techniques. Environ. Sci. Technol. 22(7)-740-745.
Marrin, D.L. and G.M. Thompson. 1984. Remote Detection
of Volatile Organic Contaminants in Ground Water Via
Shallow Soil Gas Sampling. In: Proc. IstNWWA/API
Conf. Petroleum Hydrocarbons and Organic Chemicals
in Ground Water-Prevention, Detection and Restora-
tion, National Water Well Association, Dublin, OH, pp.
172-187.
Marrin, D.L. and G.M. Thompson. 1987. Gaseous Behavior
of TCE Overlying a Contaminated Aquifer. Ground Wa-
ter 25(l):21-27.
Marsh, J.M. and J.W. Lloyd. 1980. Details of Hydrochemical
Variations in Flowing Wells. Ground Water 18(4): 366-
••> ^70
373.
Mason, B.J. 1983. Preparation of Soil Sampling Protocol:
Techniques and Strategies. EPA/600/4-83-020 (NTIS
PB83-206979).
McKee, C.R. and A.C. Bumb. 1988. A Three-Dimensional
Analytical Model to Aid in Selecting Monitoring Loca-
tions in the Vadose Zone. Ground Water Monitoring
Review 8(2):125-136.
McNichols, R.J. and C.B. Davis. 1988. Statistical Issues and
Problems in Ground Water Detection Monitoring at Haz-
ardous Waste Facilities. Ground Water Monitoring Re-
view 8(4): 135-150.
Meade, J.P. and W.D. Ellis. 1985. Decontamination Tech-
niques for Mobile Response Equipment Used at Waste
Sites (State-of-the-Art Survey). EPA/600/2-85/105 (NTIS
PB85-247021).
Mickam, J.T., R. Bellandi, and E.G. Tifft, Jr. 1989. Equipment
Decontamination Procedures for Ground Water and Va-
dose Zone Monitoring Programs: Status and Prospects.
Ground Water Monitoring Review 9(2): 100-121.
Morrison, R.D. and B. Lowery. 1990. Sampling Radius of a
Porous Cup Sampler Experimental Results. Ground Wa-
ter 28(2):262-267.
Munch, J.H. and R.W.D. Killey. 1985. Equipment and Meth-
odology for Sampling and Testing Cohesionless Sedi-
ments. Ground Water Monitoring Review 5(l):38-42.
Myers, R.G., C.W. Swallow, andD.E. Kissel. 1989. A Method
to Secure, Leach and Incubate Undisturbed Soil Cores.
Soil Sci. Soc. Am. J. 53:467-471.
National Council of the Paper Industry for Air and Stream
Improvement (NCASI). 1982. A Guide to Groundwater
Sampling. Technical Bulletin 362, NCASI, New York,
NY.
New Jersey Department of Environmental Protection (NJDEP).
1988. Field Sampling Procedures Manual. Hazardous
Waste Program, NJDEP, Trenton, NJ.
Newman, W., J.M. Armstrong, and M. Ettenhofer. 1988. An
Improved Soil Gas Survey Method Using Adsorbent
Tubes for Sample Collection. In: Proc. 2nd Nat. Outdoor
Action Conf. on Aquifer Restoration, Ground Water Moni-
toring and Geophysical Methods, National Water Well
Association, Dublin, OH, pp. 1033-1049.
Niaki, S. and J.A. Broscious. 1986. Underground Tank Leak
Detection Methods: A State-of-the-Art Review. EPA/
600/2-86/001 (NTIS PB86-137155).
Nielsen, D.M. (ed.). 1991. Practical Handbook of Ground
Water Monitoring. Lewis Publishers, Chelsea, MI (pub-
lished in cooperation with National Water Well Associa-
tion, Dublin, OH), 717 pp.
Nielsen, D.M. and G.L. Yeates. 1985. A Comparison of
Sampling Mechanisms Available for Small-Diameter
Ground Water Monitoring Wells. Ground Water Moni-
toring Review 5(2):83-99.
151
-------
Nightingale, H.I. and W.C. Bianchi. 1979. Influence of Well
Water Quality Variability on Sampling Decisions and
Monitoring. Water Resources Bulletin 15(5):1394-1407.
Nightingale, H.I. and W.C. Bianchi. 1980. Well Water Qual-
ity Changes Correlated with Well Pumping Time and
Aquifer Parameters-Fresno, California. Ground Water
18:275-280.
Olivems, J.P., D.A. Vroblesky, and M.M. Lorah. 1988. In-
creasing Purging Efficiency Through the Use of Inflat-
able Packers. In: Proc. 2nd Nat. Outdoor Action Conf. on
Aquifer Restoration, Ground Water Monitoring and Geo-
physical Methods, National Water Well Association,
Dublin, OH, pp. 457-469.
Oudjik, G. and K. Mujica. 1989. Handbook for Identification,
Location and Investigation of Pollution Sources Affect:
ing Ground Water. National Water Well Association,
Dublin, OH.
Palmer, C.D., J.F. Keely, and W. Fish. 1987. Potential for
Solute Retardation on Monitoring Well Sand Packs and
Its Effect on Purging Requirements for Ground Water
Sampling. Ground Water Monitoring Review 7(2):40-47.
Panko, A.W. and P. Earth. 1988. Chemical Stability Pnor to
Ground-Water Sampling: A Review of Current Well
Purging Methods. In: Ground-Water Contamination: Field
Methods, A.G. Collins and A.I. Johnson (eds.), ASTM
STP 963, American Society for Testing and Materials,
Philadelphia, PA, pp. 232-239
Pennino, J.D. 1988. There's No Such Things as a Representa-
tive Ground Water Sample. Ground Water Monitoring
Review 8(3):4-9.
Perry, C.A. and R.J. Hart. 1985. Installation of Observation
Wells on Hazardous Waste Sites in Kansas Using a
Hollow-Stem Auger. Ground Water Monitoring Review
5(4):70-73.
Peterson, R.G. and L.D. Calvin. 1986. Sampling. In: Methods
of Soil Analysis, Part I—Physical and Mineralogical
Methods, 2nd ed., A. Klute (ed.), ASA Monograph No. 9,
American Society of Agronomy, Madison, WI, pp. 33-
51.
Pitard, F.F. 1989. Pierre Gy's Sampling Theory and Sampling
practice, Vol. I Heterogeneity and Sampling, Vol. II
Sampling Correctness and Sampling. CRC Press, Boca
Raton, FL.
Pohhnann, K.F. and J.W. Hess. 1988. Generalized Ground
Water Sampling Device Matrix. Ground Water Monitor-
ing Review 8(4): 82-84.
Puls, R.W, and M.J. Barcelona. 1989. Ground Water Sam-
pling for Metals Analyses. Superfund Ground Water
Issue Paper. EPA/548/4-89/001.
Rajagopal, R. 1986. The Effect of Sampling Frequency on
Ground Water Quality Characterization. Ground Water
Monitoring Review 6(4)65-73.
Rajagopal, R. and L.R. Williams. 1989. Economics of Sample
Compositing as a Screening Tool in Ground Water Qual-
ity Monitoring. Ground Water Monitoring Review
9(1):186-192.
Rehm, B.W., T.R. Stolzenburg, D.G. Nichols. 1985. Field
Measurement Methods for Hydrogeologic Investigations:
A Critical Review of the Literature. EPRI AA-4301.
Electric Power Research Institute, Palo Alto, CA.
Robbins, G.A. 1989. Influence of Using Purged and Partially
Penetrating Monitoring Wells on Contaminant Detection.
Mapping, and Modeling. Ground Water 27(2):155 -162.'
Robin, M.J.L. and R.W. Gillham. 1987. Field Evaluation of
Well Purging Procedures. Ground Water Monitoring Re-
view 7 4):85-93.
Rose, S.R. and A. Long. 1988. Monitoring Dissolved Oxygen
in Ground Water Some Basic Considerations. Ground
Water Monitoring Review 8(l):93-97.
Roy W.R., S.C. Mravik, I.G. Krapac, D.R. Dickerson, and
R.A. Griggin. 1989. Geochemical Interactions of Hazard-
ous Wastes with Geological Formations in Deep-Well
Systems. ISGS Environmental Geology Notes 130. Illi-
nois State Geological Survey, Champaign, IL.
Scalf M.R., J.F. McNabb, W.J. Dunlap, R.L. Cosby, and J.
Fryberger. 1981. Manual of Ground-Water Quality Sam-
pling Procedures. EPA/600/2-81/160, (NTIS PB82-
103045). Also published in NWWA/EPA Series, National
Water Well Association, Dublin OH.
Schalla, R., D.A. Myers, M.A. Simmons, J.M. Thomas, and
A.P. Toste. 1988. The Sensitivity of Four Monitoring
Well Sampling Systems to Low Concentrations of Three
Volatile Organics. Ground Water Monitoring Review
8(3):90-96.
Schmidt, K.D. 1982. How Representative are Water Samples
Collected from Wells? In: Proc. 2nd Nat. Symp. on
Aquifer Restoration and Ground Water Monitoring, Na-
tional Water Well Association, Dublin, OH, pp. 117-128.
Sgambat, J.P. and R. Stedinger. 1981. Confidence in Ground-
Water Monitoring. Ground Water Monitoring Review
l(Spnng): 62-69.
Shangraw, T.C., D.P. Michaud, T.M. Murphy. 1988. Verifica-
tion of the Utility of a Photovac Gas Chromatography for
Conduct of Soil Gas Surveys. In: Proc. 2nd Nat. Outdoor
Action Conf. on Aquifer Restoration, Ground Water Moni-
toring and Geophysical Methods, National Water Well
Association, Dublin, OH, pp. 1089-1108.
Simmons, M.S. 1991. Hazardous Waste Measurements. Lewis
Publishers, Chelsea, MI, 315 pp.
152
-------
Sisk, S.W. 1981. NEIC Manual for Groundwater/Subsurface
Investigations at Hazardous Waste Sites. EPA/330/9-81-
002 (NTIS PB82-103755).
Slater, J.P. and F.R. McLaren. 1983. Sampling and Analysis
of Soil for Volatile Organic Compounds. In: Proc.
NWWA/EPA Conf. on Characterization and Monitoring
of the Vadose (Unsaturated) Zone (Las Vegas, NV),
National Water Well Association, Dublin, OH, pp. 623-
658.
Slawson, Jr., G.C., K.E. Kelly, andL.G. Everett. 1982. Evalu-
ation of Ground Water Pumping and Bailing Methods—
Application in the Oil Shale Industry. Ground Water
Monitoring Review 2(3):27-32.
Slonim, M.J. 1957. Sampling in a Nutshell. J. Am. Statistical
Ass. 52:143-161.
Smith, J. S., D.P. Steele, M.J. Malley, and M.A. Bryant. 1988.
Groundwater Sampling. In: Principles of Environmental
Sampling, L.H. Keith (ed.), ACS Professional Reference
Book, American Chemical Society, Washington, DC, pp.
255-260.
Sophocleous, M., J.E. Paschetto, and R.A. Olea. 1982. Ground-
Water Network Design for Northwest Kansas, Using the
Theory of Regionalized Variables. Ground Water 20:48-
58.
Spittler, T.M., W.S Clifford, and L.G. Fitch. 1985. A New
Method for Detection of Organic Vapors in the Vadose
Zone. In: Proc. 2nd Conf. on Characterization and Moni-
toring in the Vadose (Unsaturated) Zone, National Water
Well Association, Dublin, OH, pp. 236-246.
Spruill, T.B. 1988. Use of Total Organic Carbon as an Indica-
tor of Contamination from an Oil Refinery, South-Cen-
tral Kansas. Ground Water Monitoring Review 8(3)76-82.
Starks, T.H, K.W. Brown, and N.J. Fisher. 1986. Preliminary
Monitoring Design for Metal Pollution in Palmerton,
Pennsylvania. In: Quality Control in Remedial Site In-
vestigation: Hazardous and Industrial Solid Waste Test-
ing, 5th Volume, C.L. Perket (ed.), ASTM STP 925,
American Society for Testing and Materials, Philadel-
phia, PA, pp. 57-66.
Steel, T.D. 1986. Converting Water Quality Information Goals
into Statistical Design Criteria. In: Monitoring to Detect
Changes in Water Quality Series, D. Lemer (ed.), Int.
Ass. of Hydrological Sciences Pub. No. 157, pp. 71-79,
Stevenson, C.D. 1978. Simple Apparatus for Monitoring Land
Disposal Systems by Sampling Percolating Soil Waters.
Environ. Sci. Technol. 12:329-331.
Stolzenburg, T.R. and D.G. Nichols. 1985. Preliminary Re-
sults on Chemical Changes in Groundwater Samples Due
to Sampling Devices. EPRI EA-4118. Electric Power
Research Institute, Palo Alto, CA,
Sukhatme, P.V and B.V. Sukhatme. 1970. Sampling Theory
of Surveys with Applications, 2nd ed. Iowa State Univer-
sity Press, Ames, IA.
Summers, K.V and S.A. Gherini. 1987. Sampling Guidelines
for Groundwater Quality. EPRI EA-4952. Electric Power
Research Institute, Palo Alto, CA.
Sykes, A.L., R.A. McAllister, and J.B. Homolya. 1986. Sorp-
tion of Organics by Monitoring Well Construction Mate-
rials. Ground Water Monitoring Review 6(4):44-47.
Tanur, J.M., (ed.). 1978. Statistics: A Guide to the Unknown.
Holden-Day, San Francisco.
Tinlin, R.M. 1976. Monitoring Groundwater Quality: Illustra-
tive Examples. EPA/600/4-76-036 (NTIS PB257 936).
Todd, O.K., R.M. Tinlin, K.D. Schmidt, and L.G. Everett.
1976. Monitoring Ground-Water Quality: Monitoring
Methodology. EPA/600/4-76-026 (NTIS PB256-068).
UNESCO. 1983. proceedings of the Symposium-Methods
and Instrumentation for the Investigation of Groundwater
Systems. Committee for Hydrological Research, CHO-
TNO, The Hague, The Netherlands.
Unwin, J. and V. Maltby. 1988. Investigations of Techniques
for Purging Ground-Water Monitoring Wells and Sam-
pling Ground Water for Volatile Organic Compounds. In:
Ground-Water Contamination: Field Methods, A.G.
Collins and A.I. Johnson (eds.), ASTM STP 963, Americ-
an Society for Testing and Materials, Philadelphia, PA,
pp. 240-252.
U.S. Environmental Protection Agency (EPA). 1985. RCRA
Ground-Water Monitoring Compliance Order Guidance.
EPA Office of Solid Waste and Emergency Response
(NTIS PB87-193710).
U.S. Environmental Protection Agency (EPA). 1986a. RCRA
Ground Water Monitoring Technical Enforcement Guid-
ance Document. EPA OSWER-9950.1. Also published in
NWWA/EPA Series, National Water Well Association,
Dublin, OH.
US. Environmental Protection Agency (EPA). 1986b. Test
Methods for Evaluating Solid Waste, 3rd ed., Vol. II
Field Manual Physical/Chemical Methods. EPA/530/
SW-846 (NTIS PB88-239223) First update, 3rd ed. EPA/
530/SW-86.3-1 (NTIS PB89-148076).
U.S. Environmental Protection Agency (EPA). 1986c. Permit
Guidance Manual on Unsaturated Zone Monitoring for
Hazardous Waste Land Treatment Units. EPA/530/SW-
86-040.
U.S Geological Survey. 1977+. National Handbook of Rec-
ommended Methods for Water Data Acquisition. USGS
Office of Water Data Coordination, Reston, VA.
153
-------
van Duijvenbooden, W. and H.G. van Waegeningh (eds.).
1987. Vulnerability of Soil and Groundwater to Pollut-
ants. Committee for Hydrological Research, CHO-TNO,
The Hague, The Netherlands.
Voorhees, K.J., J.C. Hickey, and R.W. Klusman. 1984. Analy-
sis of Groundwater Contamination by a New Surface
Static Trapping/Mass Spectrometry Technique. Anal.
Chem. 56:2602-2604.
Williams, L.R., R.W. Leggett, M.L. Espegren, and C.A. Little.
1989. Optimization of Sampling for the Determination of
Mean Radium-226 Concentration in Surface Soil. Envi-
ronmental Monitoring and Assessment 12:83-96.
Williams, W.H. 1978. A Sampler on Sampling. John Wiley&
Sons, New York.
Wilson, L.G. 1980. Monitoring in the Vadose Zone: A Re-
view of Technical Elements. EPA/600/7-80-134 (NTIS
PB81-125817).
Wilson, L.G. 1981. Monitoring in the Vadose Zone: Part I.
Ground Water Monitoring Review 1(3): 32-41.
Wilson, L,G. 1982. Monitoring in the Vadose Zonti Part II.
Ground Water Monitoring Review 2(1 ):3 1-42.
Wilson, L.G. 1983. Monitoring in the Vadose Zone Part III.
Ground Water Monitoring Review 3(2): 155-166.
Wittmann, S. G., K.J. Qumn, and R.D. Lee. 1985. Use of Soil
Gas Sampling Techniques for Assessment of Ground
Water Contamination. In: Proc. NWWA/API Conf. Pe-
troleum Hydrocarbons and Organic Chemicals in Ground
Water—Prevention, Detection and Restoration, 1985,
National Water Well Association, Dublin, OH, pp. 291-
309.
Wood, W .W. 1976. Guidelines for Collection and Field Analy-
sis of Groundwater Samples for Selected Unstable Con-
stituents. U.S. Geological Survey TWI 1-D2.
Yates, F. 1980. Sampling Methods for Censuses and Surveys,
4th ed. MacMillan, New York.
Zapico, M.M., W. Vales, and J.A. Cherry. 1987. A Wireline
Piston Core Barrel for Sampling Cohesionless Sand and
Gravel Below the Water Table. Ground Water Monitor-
ing Review 7(3); 74-82.
154
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PART II: PHYSICAL AND CHEMICAL PROCESSES IN THE SUBSURFACE
Chapter 10
Physiochemical Processes: Organic Contaminants
Carl D. Palmer and Richard L. Johnson
10.1 Overview of Physiochemical Processes
The characterization of hazardous waste sites to design
remediation strategies requires a broad range of background
information. As discussed in Chapter 9, good sampling meth-
ods and strategies are required to determine the contamination
level and the extent to which contaminants have moved within
the subsurface. Understanding of the physical processes dis-
cussed in Chapters 4 and 5 allows determination of the rate
and direction in which contaminated ground water is flowing.
This information also can be used to determine whether the
contaminants will be captured and removed by extraction
wells. However, this information by itself is not sufficient for
optimal choice of remediation schemes. Critical questions
such as how much water must pass through a section of an
aquifer to remove the contaminants or how much time is
required for contaminants to diffuse out of low-permeability
zones also must be answered. The answers to these questions
depends on the physiochemical processes occurring within
the subsurface.
The next three chapters address the physiochemical
processes that recur within the subsurface, the parameters
required for their characterization, and the implications of
these processes for remediation design. In this chapter, the
discussion is limited to processes occurring below the water
table that affect the concentration, transport, and hence re-
moval of organic contaminants. Chapter 11 addresses the
transport of volatile organic compounds through the unsatur-
ated zone, and Chapter 12 discusses inorganic contaminants.
The design of optimal remediation schemes often re-
quires some "prediction" of the distribution of contaminants
within the subsurface over time. These predictions then can be
used to evaluate different remediation scenarios. The basis for
making such predictions is generally the application of the
concepts of mass balance. A common method for applying
mass balance concepts to dissolved chemical constituents in
ground-water systems is the advection-dispersion equation,
which is written in its one-dimensioml form as:
[10-1]
-v-± RXN
3x2 3x dt
where v is the ground-water velocity (L/T), D is the dispersion
coefficient (L2/T), C is the concentration of the dissolved
constituent (M/L3), t is time, and RXN represents a general
chemical reaction term. The frost term in eq. 10-1 describes
the net advective flux of the contaminant in and out of a
volume of the aquifer (Figure 10-1). The second term de-
scribes the net dispersive flux of the contaminant. The first
term on the right-hand side of the equation describes the
change in concentration of the contaminant in the water
contained within the volume of aquifer. The second term on
the right-hand side represents the amount of contaminant that
may be added or lost to the ground water by some chemical or
biological reaction. If there is no reaction term, then the
equation describes the transport of a conservative, nonreacting
tracer such as chloride or bromide. More detailed information
about the development and derivation of eq. 10-1 is found in
Palmer and Johnson (1989), Gillham and Cherry (1982),
Freeze and Cherry (1979), or Bear (1979, 1969).
Some understanding of this mass balance equation is
useful even to the individual who is not directly responsible
for making mathematical representations of the distribution of
contaminants within the subsurface. The equation is an ex-
ample of the current understanding of the processes control-
ling the fate and transport of contaminants in the subsurface.
The equation lists the parameters that should be quantified
either by performing appropriate field or laboratory measure-
ments or by using the best known values. The results of the
application of this modeling are unlikely to ever exactly
"predict" how the contaminants behave at a particular field
site but they can provide a general set of expectations that are
useful in the design of a remedial system. These results also
can be used to compare aquifer remediation performance.
According to eq. 10-1, two parameters that must be
determined are the ground-water velocity, v, and the disper-
sion coefficient, D. These parameters are described in Chap
ters 4 and 5 as well as in other sources (e.g., Palmer and
Johnson, 1989 a,b). Chemical processes that can affect the
fate and transport of organic contaminants below the water
table include (1) abiotic degradation, (2) biotic degradation,
(3) dissolution nonaqueous phase liquids (NAPLs), (4) sorp-
tion reactions, and (5) ionization. Both abiotic and biotic
155
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Transport of Reactive Solutes
Flux In
~3C
Advective Dispersive
Flux Flux
Flux Out
dC ndC
Advective
Flux
Mass Balance Equation
_ O C oC oC nv»i
D —!, -—_ = —.. ±Wf/v
9x^ ox df
Dispersive Advective Change in peaction
Term Term Mass per -rsfm
Unit Time
Dispersive
Flux
Figure 10-1. Mass balance equation for the transport of reactive solutes through porous media.
degradation are discussed in Chapter 13. The discussion in
this chapter is limited to the three latter processes.
10.2 Dissolution of Nonaqueous Phase Liquids
Many of the organic chemicals of environmental concern
enter the subsurface in the nonaqueous phase. How these
solvents move through the soil depends on the grain size of
the aquifer material, the degree of water saturation in the pore
space, and the density and viscosity of the solvent relative to
water (Palmer and Johnson, 1989c; Schwille, 1988). For
example, if there is a spill of nonaqueous phase liquid that has
a density greater than water (DNAPL), as it flows through the
unsaturated zone, because the water is in the wetting phase, it
will pass through the center of the pores. If there is residual
water within the unsaturated zone then the combination of
higher density and lower viscosity of the DNAPL relative to
water results in unstable flow or significant fingering of the
DNAPL as it moves through the porous media. If the spill is
large enough so that the DNAPL can penetrate the capillary
fringe and move below the water table, this fingering contin-
ues to occur. The transport of the DNAPLs is also very
sensitive to small changes in permeability. Therefore, the
DNAPL tends to spread laterally as it encounters lenses of
finer grained material in the subsurface. This combination of
viscous fingering and lateral flow results in a series of fingers
and pools of DNAPL. The DNAPL in the fingers tends to
drain to some residual saturation while the pools contain
DNAPL above the residual saturation.
As ground water flows through the fingers, the DNAPL is
dissolved by the passing ground water. Laboratory experi-
ments (Anderson, 1988; Anderson et al, 1987) using a 15-cm-
diameter cylindrical finger of tetrachloroethylene (TeCE)
(Figure 10-2) demonstrate that the ground water passing
through the fingers can quickly reach saturation with the
TeCE. This was found to be true for ground-water velocities
ranging from 10 to 100 cm/day (Figure 10-3). However, these
results do not imply that where a DNAPL spill has occurred
the sampled ground water is saturated with the solvent. In-
deed, sampling results usually indicate that most waters are
highly undersaturated with respect to the DNAPLs. Although
Cylinder of TeCE at
Residual Saturation
Experimental Sand Tank
1m x 1m x 1m
Figure 10-2. Cylindrical source of tetrachloroethylene (TeCE)
used in the experiments by Anderson (1988).
the water that passes through the fingers or very close to the
pools of DNAPL within the subsurface is saturated with the
DNAPL, mass transfer of the dissolved DNAPL to the areas
further from these fingers and pools is predominantly by
molecular diffusion. As a result, many areas within the aquifer
that lie between the pools and fingers contain little or no
dissolved solvent. While the distance between such fingers
and pools is generally unknown, it is probably at least as great
as the mean distance between the small-scale beds within the
aquifer. For the Borden aquifer in Ontario, Sudicky (1986)
found this distance to be about 10 cm in the vertical direction.
A typical monitoring well would have an intake length of at
least 2 m. Thus, the water saturated with the solvent is mixed
156
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10 cm/day
30 cm/day
100 cm/day
200
160
120\-
ad
40-
-20 -16 -12 -8 -4 0 4 8 12 16 20
Distance from Plume Center (cm)
10-3. Concentration of TeCE across the flow field at the end tank in the sand box experiments conducted byAnderson
(1988).
with the uncontaminated ground water resulting in measured
concentrations that are substantially below saturation.
Estimating the time required to remove the nonaqueous
phase liquid from the subsurface is difficult. Estimates require
knowledge of the amount that was spilled and the distribution
of the solvent within the aquifer. While the former piece of
information is often difficult to obtain, the latter is virtually
impossible. If the solvent is assumed to be uniformly distrib-
uted (a residual saturation, SJ within the aquifer, and the
ground water flowing through the aquifer instantaneously
equilibrates with the solvent, then the time required to remove
the solvent by dissolution, tr, is
tr = S0L/(C q)
[10-2]
same velocity as the ground water but can be slowed by their
interaction with the soil matrix. This interaction with the soil
is often described graphically as an adsorption isotherm. An
adsorption isotherm is simply a plot of the concentration of
the contaminant on the soil versus the concentration of the
contaminant in solution. Isotherms are so named because they
are conducted at constant temperature. Different types of
adsorption isotherms are defined according to their general
shape and mathematical representation. For a Langmuir iso-
therm, the concentrations on the soil increase with increasing
ground-water concentrations until a maximum concentration
on the soil is reached (Figure 10-4). The isotherm can be
represented by the equation
[10-3]
where q is the porosity of the aquifer, L is the length of the
aquifer containing the solvent through which the ground water
flows, Ccqis the equilibrium concentration of the contaminant
in the ground water, and q is the ground-water flux. Estimates
of removal times based on eq. 10-2, however, underestimate
the actual removal time because the equation does not account
for the role of soil heterogeneity, the differential times the
ground water takes to flow along different flowlines, or the
limitations in mass transfer of pools of NAPL that are above
residual saturation. If a pump-and-treat remediation scheme is
already in place, remediation time can be roughly estimated
by dividing the total mass of solvent in the aquifer by the mass
being removed per unit time by extraction wells.
10.3 Sorption Phenomena
10.3.1 Adsorption Isotherms
Once an organic compound has been dissolved into the
ground water, it will be transported away from the source area
by ground-water flow. The contaminants do not travel at the
SMAX
Aqueous Concentration
Figure 10-4. Lndmuir adsorption isotherm.
157
-------
a>
Solution Concentration
Figure 10-5. Freundlich adsorption isotherm.
where S (M/M) is the concentration on the soil, Smn (M/M) is
the maximum concentration on the soil, K(L3/M) is the
Langmuir adsorption constant, and C (M/L3) is the concentra-
tion in the ground water. A Freundlich (or Kiister) isotherm is
given by the equation:
= KC"
[104]
where K is the Freundlich adsorption constant and a is a
positive parameter. The shape of a Freundlich isotherm de-
pends on the value of a. If a is greater than 1.0, the isotherm
becomes steeper with increasing concentrations in the ground
water. If a is less than 1.0, the isotherm becomes steeper at
lower concentrations (Figure 10-5).
A linear isotherm is a special case of the Freundlich
isotherm where the parameter a is equal to unity. Linear
isotherms are of particular interest because (1) many nonpo-
lar, hydrophobic organic compounds tend to follow linear
isotherms (Figure 10-6) over a wide range of conditions and
(2) the application of a linear isotherm simplifies the math-
ematical model used to simulate the rate of contaminant
movement in the subsurface and reduces the number of pa-
rameters that need to be obtained during characterization.
Another way of representing the partitioning between the
soil and the ground water is by a "partition coefficient," Kp.
The partition coefficient is the ratio of the change in concen-
tration of the contaminant on the soil to the change in concen-
tration of the contaminant in the ground water or more simply,
the slope of the isotherm. When the isotherm for a particular
soil is linear, the partition coefficient is constant.
1200
7,1,1-Trichloroethane
^ 1,1,2,2-
Tetrachloroethane
1,2-Dichloroethane
0 400 800 1200 1600 2000 2400
Aqueous Concentration frig/i)
Figure 10-6. Linear sorption isotherms obtained for several priority pollutants (after Chlou et al., 1979).
158
-------
The partition coefficient of an organic chemical is not
constant for every soil. In general, Kpincreases as the fraction
of organic carbon, foc, increases in the soil (Karickhoff, 1981).
In other words, the sorption of nonpolar, hydrophobic organic
compounds in soils is primarily an equilibrium partitioning
process into soil organic matter. Kpcan be represented by
K, = focKoc [10-5]
where Koc is the slope of the experimentally determined Kp
versus foccurves like those in Figure 10-7. Alternatively, Koc:
can be considered to be the partition coefficient for the
organic compound into an hypothetical pure organic carbon
phase.
If sorption is the primary reaction occurring in the subsur-
face, the right-hand side of eq. 10-1 represents the change in
the total mass of contaminant within a volume of the aquifer.
The total change in mass in the volume of the aquifer is equal
to the change in mass in the ground water plus the change in
mass on the solid phase. The reaction term in eq. 10-1 is then
written as (pJQ) dS/dt where pb and 0are the dry bulk density
and volumetric water content of the soil, respectively. Substi-
tuting
as
[10-6]
into this reaction term and recognizing that 9S/3C is equal to
Kpfor a linear adsorption isotherm, eq. 10-1 can now be
written as
R
R dx
[10-7]
1800
I
g
1
0.0 .005 .010 .015 .020 .025
Fraction Organic Carbon
Figure 10-7. Partition coefficients for pyrene and phenanthrene
versus the fraction of organic carbon in the soil
(after Karickhoff, 1981).
where the constant
R = 1 + Kpp/0
[10-8]
is known as the "retardation factor." The general form of the
equation only changes by the constant R. All of the math-
ematical solutions that are used to solve the transport of
nonreacting tracers can be used to solve for the transport of
nonpolar hydrophobic organic compounds if the ground-wa-
ter velocity and dispersion coefficient are divided by R.
The retardation factor can be interpreted in slightly dif-
ferent but equally valid ways. It is the ratio of the ground-
water velocity, v, to the solute velocity, vs, (i.e., R = v/vs). It is
also the ratio of the time for the solute to travel from a source
to an observation point divided by the time for the ground
water to travel that same path. The retardation factor also can
be thought to represent the number of pore volumes that must
be flushed through a soil to remove the contaminant. All of
these definitions assume that the only process occurring is
linear sorption.
Application of the new expression (eq. 10-7) requires
knowledge of the additional parameter R. This parameter can
be obtained by several methods including (1) calculation from
eq. 10-8, where Kpis obtained from correlation techniques;
(2) calculation from eq. 10-8, with Kp, obtained from batch
sorption tests; (3) measurement from column tests and (4)
estimation from field data. The other parameters in eq. 10-8
(porosity and dry bulk density) are physical parameters that
can be obtained using common techniques (see Chapter 4 and
Palmer and Johnson, 1989c).
10.3.2 Determining Retardation Factors Using
focandKoc
The relationship between the Koc value and other known
properties of organic contaminants has been examined by
numerous researchers (Kenaga and Goring, 1980; Karickhoff,
1981, Schwartzbach and Westall, 1981; Chiou et al, 1982 and
1983), For example, some research has revealed linear rela-
tionships between the log of the volubility of the contaminant
and the log (Koc) (Figure 10-8). Similarly, Karickhoff sug-
gested that the partitioning of organic contaminants into soil
organic matter must be analogous to the partitioning of those
contaminants into other organic compounds such as octanol.
He found a linear relationship (Figure 10-9) between log (KJ
and log (Kow), where Kow is the octanol-water partition coeffi-
cient. Several regression equations relating the properties of
organic compounds to the Kochave been derived (Table 10-1).
Thus, by knowing the name of the compound of interest, these
properties can be found in tables of chemical properties
(Mabey et al., 1982) and the regression equations used to
approximate Koc. The goal, however, is to determine the
partition coefficient and ultimately the retardation factor. To
do this, eq. 10-5 must be applied and a measurement of the
fraction of organic carbon must be obtained.
The many methods of measuring the amount of organic
carbon in the soil can be broadly classified as either wet
combustion or dry combustion techniques. Wet combustion
techniques involve the addition of a strong oxidizing agent
159
-------
7
6
5
4
3
2
1
0
-1
Log Kx = -0.55 log S +3.64
(S in mg/L)
-4-3-2-101234567
Log S (mg/L)
Figure 10-8. Log Kot versus logarithm of the volubility of the compound in water (after Kenaga and Goring, 1980).
Figure 10-9. Log K0 versus the octanol-water partition coefficient. Data from Karickhoff (1981).
such as bichromate to the soil. There are several such wet
combustion techniques including the Walkley-Black method
and the modified Mebius procedure; these procedures are
discussed in detail by Nelson and Sommers (1982). In spite of
some limitations, these methods can provide a relatively rapid
and inexpensive method for obtaining estimates of foc.
Dry combustion methods generally involve heating the
soil sample in the presence of oxygen. The oxygen reacts with
the soil carbon to form carbon dioxide that can be detected by
a variety of techniques.
To estimate the linear retardation factor, the Koc obtained
from one (or more) of the regression equations given in Table
10-1 is multiplied by the fraction of organic carbon to yield
the partition coefficient (eq. 10-1). The retardation factor is
obtained from the Kp, pb, and d by eq. 10-8.
There are several limitations to the use of the correlation
techniques described above. The linear relationship between
focand Kpis not always easy to determine. In particular, the
relationship is most likely to fail when (1) the focis very low
(<0,001), (2) when there are large amounts of swelling clays
present, and (3) the organic compound is polar (e.g., com-
pounds that contain amine or carboxylic acid groups) (Pankow,
160
-------
Table 10-1. Some Reported Correlation Equations
Equation
Data Base1
Reference
log KK * 0.544 log KM + 1.377
log KK = 1.00 log Kw- 0.21
log Kx = -0.55 log Sw + 3.64 »
log Kx = -0.56 log Sw + 0.93 •*
log KK= -0.54 log xa + 0.44 "
aromatic hydrocarbons (8)
carboxylic acids and esters (5)
P containing insecticides (5)
ureas and uacils (7)
symmetrical triazines (6)
miscellaneous (14)
polycyclic aromatics (8)
chlorinated hydrocarbons (2)
aromatic hydrocarbons (8)
carboxylic acids and esters (5)
P containing insecticides (5)
ureas and uacils (7)
symmetrical triazines (6)
miscellaneous (14)
polychlorinated biphenyls (3)
pesticides (4)
halogenated ethanes & propanes (6)
tetrachloroethene
1,2-dichiorobenzene
polycyclic aromatics
chlorinated hydrocarbons
Kenaga and Goring (1978)
Karickhoffetal. (1979)
Kenaga and Goring (1978)
Chiouetal.(1979)
Karickhoff (1979)
'Number in parentheses refer to the number of compounds in data base.
'Swis the solubility of the compound in water in ppm.
'Derived from the original equation assuming Ka = 1.7 Kom
"XJs the mole fraction solubulity at 25°C.
After Pankow, 1984
1984). There are also several reasons why the relationship
between log (Kocand log (KoW) may not always be linear
(Pankow, 1984). If mechanisms other than simple partitioning
into soil organic carbon are contributing to the adsorption of
the organic contaminant, then the Koc value, computed as the
ratio K/foc, will be in error. Also, If the molecule is large it
may not fit into the soil organic matter to the same extent as it
would in octanol (steric limitations). Finally, if the adsorption
is strong, a contaminant may take a substantial period of time
to equilibrate with the soil organic carbon.
10.3.3 Determining Retardation Factors Using
Batch Tests
Retardation factors also can be measured with batch tests.
These tests are, in principle, easy to perform, and the method
is outlined in Figure 10-10. A known volume of solution, Vw,
containing an initial concentration, C0, of a contaminant is
placed into a container. A known mass of soil, Ms, is then
added and the mixture is shaken and allowed to equilibrate.
The soil then is separated from the solution by centrifuging,
and an aliquot of the supernatant is sampled. The concentra-
tion of the contaminant in this aliquot, C, is measured and the
concentration on the soil, S, is calculated by
Batch Adsorption Tests
Solution with
Contaminant
Soil with
Organic Matter
Shake and
Equilibrate
S=VW(C0-C)/MS
S = Vw(C0-C)/M
[10-9]
Sample and Measure
Contaminant Concentration in
Solution
Figure 10-10. Batch adsorption tests.
161
-------
This test can be run several times with different initial
concentrations or different masses of soil. The result is a
series of contaminant concentrations with corresponding aque-
ous phase concentrations that yield an isotherm when they are
plotted. If the isotherm is linear, the slope, or partition coeffi-
cient, can be easily determined. The retardation factor then
cart be calculated from Kp,/>4, and 0 using eq. 10-8.
Prior to conducting such batch adsorption tests, the soil is
prepared by drying and then sieving through a 2-mm sieve.
The sieving is to ensure that aggregated soil particles are
relatively small, thus reducing the time for the contaminant to
diffuse into the particles and equilibrate with the soil. Another
important preparatory step is to estimate the Kp using, for
example, the correlation methods described in Section 10.3.2.
This is important in choosing the proper amount of soil to use
in the tests. If Kpis large and too much soil is added to the
reaction vessel, then most of the contaminant is partitioned to
the soil and the concentration in solution cannot be accurately
determined. Similarly, if Kpis small and too little soil is added
to the reaction vessel, then the measured contaminant concen-
tration falls within the analytical error of the initial concentra-
tion and an accurate estimate of the contaminant concentration
of the soil cannot be obtained. Both of these cases lead to poor
measures of the partition coefficient.
There are some problems that complicate the use of batch
tests for determining Kp. For example, batch tests assume that
equilibrium is established between the soil and the solution,
but some contaminants may take a very long period of time to
equilibrate. Experiments on the resorption of hexachloro-
benzene from soils (Karickhoff and Morris, 1985) indicated
that even after 35 days equilibrium was not obtained (Figure
10-11).
Another problem involves nonsettling particles. The sepa-
ration of the soil and the water is assumed to be complete
before sampling of the supernatural however, very fine, col-
loidal-size particles may remain in suspension. The contami-
nants attached, to these particles are stripped during the analysis
of the water, which causes overestimation of the aqueous
phase concentration. This results in underestimation of the
partition coefficient (e.g., Gschwend and Wu, 1984). The
magnitude of the effect depends on the concentration of
nonsettling particles (NSPS) and the true partition coefficient
onto those particles (Figure 10- 12). If the partition coefficient
is small, then most of the mass of the contaminant is in
solution and the error caused by the NSPS is negligible. If the
partition coefficient is large, then a significant mass of the
contaminant is really partitioned onto the soil particles caus-
ing significant errors in the aqueous phase concentration and
hence C
P
A third problem arises from the loss of contaminant by
volatilization during equilibration, sampling, and analysis.
This problem can be minimized by eliminating head-space
and using properly sealed reaction vessels.
Uncontaminated background soils are recommended for
batch adsorption tests. If the soils contain any NAPLs, the
contaminant being investigated will partition into the NAPL,
yielding a potentially large and incorrect partition coefficient.
Once Kpis determined in the batch test, the retardation factor,
R, can be estimated by using eq. 10-8.
10.3.4 Determining Retardation Factors from
Column Tests
A third method for estimating linear repartition factors is
with column tests. In these tests, a column of soil is prepared,
0.8
0.6
0.4
0.2
Sediment 13
3 x 10-3g/mL
Sediment 4
5x10-3g/mL
10
20
Time (Davs)
30
40
Figure 10-11. The fraction of hexachlorobenzene sorbed to two soils versus time during desorption teats (after Karickhoff and
Morris, 1985).
162
-------
1x10*
3x105
lx10
3x10'
1X104
3x10*
1x10"
Kp = 1x10c
K0=3x10'1
,= 1x104
Kp=3x103
0.1 0.2 0.5 1.0 2.0 5.0 10.0
Concentration of Nonsettling Particles (mg/L)
Figure 10-12. The effects of nonsettling particles on the
observed partition coefficient (after Pankow,
1984).
and a solution containing a nonadsorbing tracer and the con-
taminant of interest is run through the column (Figure 10-13).
The concentrations of the tracer and contaminant can be
measured in the water that has passed through the column.
The retardation factor is then the ratio of the time (or volume)
for the center of mass of the contaminant to break through the
column to the time (or volume) for the center of mass of the
nonreactive tracer to break through the column. This tech-
nique provides a direct measure of R; however, it is only
well suited for those contaminants that have a relatively
small (< 10) retardation factor. Contaminants with retardation
factors much greater than 10 require too much time to mea-
sure to be practical. Other disadvantages of using column tests
include the slow flow rates in fine-grained material, the de-
struction of soil structure by soil repacking, and the difficulty
in distinguishing kinetic behavior from the heterogeneous
packing within the column.
10.3.5 Determining Retardation Factors from
Field Data
Site-specific field information obtained during the Reme-
dial Investigation/Feasibility Study (RI/FS) can, in some cases,
be used to estimate contaminant retardation. While in prin-
ciple retardation factors can be back-calculated from break-
through curves obtained at monitoring wells or through the
spatial distribution of the contaminants in the subsurface, in
practice, only the latter is likely to be obtained. The retarda-
tion factors can be estimated by dividing the velocity of
ground water by the velocity of the contaminant. The ground-
water velocity can be estimated from Darcy's Law and the
porosity, or alternatively by the distance some nonadsorbing
solute travels after the release. The solute velocity can be
r
Water In
Water Plus
Compound
Y
t
Water Plus
Compound Out
1
oncentra
o
Non-Sorbing
*
V1
Volume
Sorting
t
A
V2
— ^~
Figure 10-13. Column tests for determination of retardation
estimated by dividing the mean distance the contaminant has
traveled by the time since its release into the subsurface. One
of the potential disadvantages of this method is that other
processes that are not included in the data analysis are occur-
ring within the aquifer. Ignoring these processes can result in
poor estimates of the retardation factor.
10.3.6 Comparison of Methods for Estimation of
Retardation
Each of the methods for estimating the retardation factor
has advantages and disadvantages. One of the key questions,
however, is how do these different methods for estimating
retardation compare. The best technique for comparison is to
look to large-scale field tracer experiments where very accu-
rate field values have been obtained. This has been done for
the Stanford-Waterloo tracer experiment that was conducted
in the sandy aquifer on Canadian Forces Base Borden in
Ontario, Canada. Details of the experiment and analysis of the
results can be found in Mackay et al. (1986); Roberts et al.
(1986); Curtis et al. (1986); Freyburg (1986); and Sudicky
(1986).
A summary of the retardation factors obtained for five
different compounds using a correlation method, batch tests,
and temporal and spatial data from the field experiment is
given in Table 10-2. The batch tests agree closely with the
field data. The correlation technique tends to consistently
underestimate the retardation factors. The underestimation of
the retardation factors may be the result of poor estimates of
the fraction of organic carbon (e.g., Powell et al., 1989) or
errors in the assumptions in eq. 10-5, or they may be the result
163
-------
Table 10.2 Comparison of Methods for Retardation Factors
Field Values
Office
Solute Estimated
CTET
BROMO
TeTE
DCB
HCB
1.3
1.2
1.3
2.3
2.3
Lab
Batch
1.9
2.0
3.6
6.9
5.4
Temporal
2.7
1.7
3.3
2.7
4.0
Spatial
2.1
2.2
4.3
6.2
6.5
After Curtis etal. (1986)
of the inherent error in the regression equations. Recall that
the regression equations are based on the logarithms of the
values; therefore, the best estimates of the Kocand hence the
retardation factor may be a factor of 2 or 3 from the "true"
value. Nonetheless, the correlation techniques do provide the
correct order of magnitude estimate of the retardation factor at
very little expense. Such values would be appropriate for the
preliminary design of the remedial strategies. If more accurate
estimates are required, then the more expensive batch or
column tests should be used. Enough samples should be
tested, however, to estimate the uncertainty of the retardation
factor for each of the important geologic units.
10.3.7 Applicability and Limitations of Linear
Partitioning and Retardation
Most of the emphasis in this section has been on the
linear partitioning and retardation model for the adsorption of
neutral, hydrophobic organic compounds in the environment.
While this model is adequate for many situations, it is impor-
tant to recognize the limitations in the assumptions so that it is
not applied to situations where it is inappropriate.
Retardation only describes the process of the partitioning
of the contaminant between the ground water and soil organic
matter. If the nonaqueous solvent phase is dissolving or the
organic compounds are degrading, then these additional pro-
cesses also must be taken into account. However, for describ-
ing the partitioning process, the linear retardation model is
reasonable for many compounds if the concentration of the
contaminant is less than 105 molar or less than half the
volubility, whichever is lower (Karickhoff et al., 1979;
Karickhoff, 1984). At high or low concentrations the linear
isotherm may deviate. Some data on the adsorption of TCE to
glacial till suggest that the partition coefficient is not constant
but may vary by as much as 50-fold over range in ground-
water concentrations from 10 to 10,000 parts per billion (ppb)
(Figure 10-14). This variation occurs even though the parti-
tion coefficient is approximately constant over the range from
100 to several thousand ppb.
The linear retardation model assumes that equilibrium is
achieved quickly. In some circumstances, the rate of adsorp-
tion and resorption can bean important factor. As mentioned
in Section 10.3.3, Karickhoff and Morris (1985) found that
during the resorption of hexachlorobenzene, equilibrium was
not achieved even after 35 days of reaction time (Figure 10-
11).
10.4 lonization and Cosolvation
Another important reaction that can affect sorption and
hence the rate of removal of organic contaminants from the
subsurface is ionization. Acidic compounds such as phenols,
catechols, quinoline, and organic acids can lose or gain pro-
tons (Ff) depending upon the pH. The resultant ions are much
more soluble and less hydrophobic than the uncharged forms.
Therefore, the ionized forms have much lower Koc values than
the uncharged forms, The pH at which this reduction in Koc
becomes substantial can be predicted based on the acidity of
the compound. This acidity is often represented as the pK,of
the compound, which is the pH at which 50 percent of the
molecules are ionized.
Table 10-3 lists pK,'s for a number of environmentally
significant ionizing compounds. For example, trichlorophenol
ionizes to a phenolate (Figure 10-15). The trichlorophenol has
a relatively large Kocvalue (2330) and readily partitions into
the soil organic matter. The ionized form is not as hydropho-
bic and its Kocvalue is substantially smaller than the Kocof the
trichlorophenol. As the pH increases, the fraction of the
phenol that is ionized increases and the Koc decreases (Figure
10-16). Therefore, the Kocvalue based on the total concentra-
tion of the phenolic compound is dependent on the degree of
ionization of the compound. While the phenolate compound
may be retarded mainly by anion adsorption to oxide surfaces
in low carbon soils, there is evidence that the phenolate also
partitions into the soil organic carbon Schellenberg et al.,
1984).
Studies with other compounds also have indicated the
relative importance of ionization of organic compounds. Stud-
ies of quinoline in low carbon soils suggest that the main
mechanism for sorption is primarily by ion adsorption (Zachara
et al., 1986; Amsworth et al., 1987).
It often is assumed that water at hazardous waste sites has
about the same chemical properties as pure water and that the
solubilities of hydrophobic organic contaminants are rela-
tively constant within a very narrow range. However, many of
the chemical properties of mixtures of solvents, such as water
and methanol, can change as the fraction of the cosolvent in
the mixture changes. The thermodynamic basis for some of
these cosolvation effects is described by Rao et al. (1985) and
Woodburn et al. (1986). Of particular interest is that the
volubility of many organic compounds can be increased by
orders of magnitude within mixtures of water and other mis-
cible solvents (Nkedi-Kizza et al., 1985; Fu and Luthy, 1986a
and 1986b; Zachara et al., 1988). For example, the partition
coefficient of anthracene decreases more than an order of
magnitude as the fraction of methanol (the cosolvent) is
increased from 0 to 50 percent (Figure 10-17).
Such cosolvation effects may be either advantageous or
disadvantageous depending on the specific problem. If these
miscible liquid cosolvents have been codisposed with priority
pollutants on site and the main concern is compliance moni-
toring, then the lower partition coefficient results in higher
transport rates to the compliance boundary. If the focus,
however, is on remediation, then the cosolvation effect may
allow a technology such as pump-and-treat to be considered a
164
-------
120
2 3
Log Aqueous Concentration (PPB)
Figure 10-14. Partition coefficients for TCE on glacial till.
Table 10-3. Acid Dissociation Constants for Several Priority
Pollutants
Compound pKt
phenol , 9.89
2-chlorophenol 8.85
2,4-dichlorophenol 7.85
2,4,6-trichlorophenol 5.99
pentachlorophenol • 4.74
2-nitrophenol 8.28
4-nitrophenol 7.15
2,4-nitrophenol 3.96
2,4-dimethyiphenol 10.6
4,6-dinitrocresol 4.35
benzidine 4.66,3.57
Source: Mabey et a/., 1982.
viable option. Alternatively, the addition of cosolvents to the
subsurface for the express purpose of enhancing the removal
of these organic contaminants in a timely and cost-effective
manner may be a possibility; however, such technology has
yet to be demonstrated in the field.
10.5 Expressions for Other Chemical Processes
The emphasis in the discussion above centered mostly on
the dissolution of the NAPL phases and equilibrium adsorp-
tion with linear partitioning. These processes are emphasized
because under many conditions they are the more important
processes controlling the rate of transport and removal of
Figure 10-15. lonization of trichlorophenol to trichlorphenolate.
organic contaminants from the subsurface. However, other
chemical processes may be taking place within the subsurface
and equilibrium may not always be a reasonable assumption.
These other equilibrium and nonequilibrium processes also
can be represented in the general expression given by eq. 10-
1. A few of the expressions for different chemical processes
are given in Table 10-4. If one of these other expressions is
required to describe the reactions that are occurring within the
subsurface, then other parameters must be measured or esti-
mated. For example, if adsorption/desorption for a particular
compound is rate-controlled rather than equilibrium-controlled,
then the rates of adsorption and resorption should be deter-
mined. These rates can be inferred from batch or column tests
similar to those described above, but they require measure-
ments over time and a more sophisticated level of interpreta-
tion and analysis. Such models should be called upon if
required for understanding the processes at a particular site.
165
-------
100 -
Figure 10-16. Percent of ionization of three different chlorophenolic compounds versus pH. Based on data from Schellenberg
et al. (1984).
1000
0.0 0.1 0.2 0.3 0.4
Fraction Co-Solvent
(Methanol)
0,5
Table 10-4. Reaction Terms for Various Chemical Processes
Process
Zero Order Production
First Order Decay
nth Order Decay
Langmuir Adsorption
Freundlich Isotherm
First Order Kinetics
Langmuir Kinetics
Nonlinear Kinectics
Reaction Term in
Mass Balance Equation
K
-KC
-KC"
(k/ejs - (k, /gjc
Figure IO-17. Partition coefficient of anthracene on three
different soils versus fraction of methanol present
as a cosolvent (adapted from Nkedi-Kizza et al.,
1985).
166
-------
10.6 References
Ainsworth, C.C., J.M. Zachara, and R.L. Schmidt. 1987.
Quinoline Sorption on Na-Montmorillonite: Contribu-
tions of the Protonated and Neutral Species. Clays and
Clay Minerals 35:121-128.
Anderson, M.A. 1988. Dissolution of Tetrachloroethylene
into Ground Water. PhD Dissertation, Oregon Graduate
Center, Beaverton, OR.
Anderson, M.A., J.F. Pankow, and R.L. Johnson. 1987. The
Dissolution of Residual Dense Non-Aqueous Phase Liq-
uid (DNAPL) from a Saturated Porous Medium. In: Proc.
NWWA/API Conf. on Petroleum Hydrocarbons and Or-
ganic Chemicals in Ground Water-Prevention, Detec-
tion and Restoration, National Water Well Association,
Dublin, OH, pp. 409-428.
Bear, J. 1969. Hydrodynamic Dispersion. In: Flow Through
Porous Media, R.J.M. De Wiest (ed.), Academic Press,
New York, pp. 109-199.
Bear, J. 1979. Hydraulics of Groundwater. McGraw-Hill,
New York.
Chiou, C.T., D.W. Schmedding and M. Manes. 1982. Parti-
tioning of Organic Compounds on Octanol-Water Sys-
tems. Environ. Sci. Technol. 16:4-10.
Chiou C.T., L.J. Peters and V.H. Freed. 1979. A Physical
Concept of Soil-Water Equilibria for Nonionic Organic
Compounds. Science 206:831-832.
Chiou, C.T., P.E. Porter and D.W. Schmedding. 1983. Parti-
tion Equilibria of Nonionic Organic Compounds Be-
tween Soil Organic Matter and Water. Environ. Sci.
Technol. 17:227-231.
Curtis, G.P, P.V. Roberts, M. Reinhard, 1986. A Natural
Gradient Experiment on Solute Transport in a Sand Aqui-
fer, 4, Sorption of Organic Solutes and Its Influence on
Mobility. Water Resources Research, 22(13):2059-2067.
Freeze, R.A. and J.A. Cherry. 1979. Groundwater. Prentice-
Hall, Englewood Cliffs, NJ.
Freyberg, D.L. 1986. A Natural Gradient Experiment on
Solute Transport in a Sand Aquifer. 2. Spatial Moments
and the Advection and Dispersion of Nonreactive Trac-
ers. Water Resources Research 22(13):2031-2046.
Fu, J.K. and R.G. Luthy. 1986a. Aromatic Compound Solu-
bility in Solvent/Water Mixtures. J. Environ. Eng. 112:328-
345.
Fu, J.K. and R.G. Luthy. 1986b. Effect of Organic Solvent on
Sorption of Aromatic Solutes onto Soils. J. Environ. Eng.
112346-366.
Gillham R.W. and J.A. Cherry. 1982. Contaminant Migration
in Saturated Unconsolidated Geologic Deposits. In: Re-
cent Trends in Hydrogeology, T.N. Narasimhan, (ed.),
Geological Society of America Special Paper 189, pp.
31-62.
Gschwend, P.M. and S-C. Wu, 1984. On the Constancy of
Sediment-Water Partition Coefficients of Hydrophobic
Organic Pollutants. Environ. Sci. Technol. 19:90-96.
Johnson, R.L., S.M. Brillante, L.M. Isabelle, J.E. Houck, and
J.F. Pankow. 1985. Migration of Chlorophenolic Com-
pounds at the Chemical Waste Disposal Site at Alkali
Lake, OR-2. Contaminant Distributions, Transport, and
Retardation. Ground Water 23(5):652-666.
Johnson, R. L, J.A. Cherry, and J.F. Pankow, 1989. Diffusive
Contaminant Transport in Natural Clay: A Field Example
and Implications for Clay-Lined Waste Disposal Sites.
Environ. Sci. Technol. 23:340-349.
Karickhoff, S.W. and K.R. Morris. 1985. Sorption Dynamics
of Hydrophobic Pollutants in Sediment Suspensions.
Environ. Toxicol. Chem. 4:469-479.
Kanckhoff, S.W., D.S. Brown and TA. Scott. 1979. Sorption
of Hydrophobic Pollutants on Natural Sediments. Water
Research 13:241-248.
Karickhoff, S.W. 1984. Organic Pollutant Sorption in Aquatic
Systems. J. Hydraulic Engineering ASCE 110:707-735.
Karickhoff, S.W. 1981. Semi-Empirical Estimation of Sorp-
tion of Hydrophobic Pollutants on Natural Sediments and
Soils. Chemosphere 10:833-846.
Kenaga, E.E. and C.A.I. Goring, 1980. Relationship between
Water Volubility, Soil-Sorption, Octanol-Water Partition-
ing, and Bioconcentration of Chemicals in Biota. In:
Aquatic Toxicology (Proc. 3rd Annual Symp. on Aquatic
Toxicology), ASTM STP 707, American Society for
Testing and Materials, Philadelphia, PA, pp. 78-115.
Mabey, W.R., et al. 1982. Aquatic Fate Process Data for
Organic Priority Pollutants. EPA/440/4-81-014 (NTIS
PB87-169090), Chapter 4,
Mackay, D. and B. Powers. 1987. Sorption of Hydrophobic
Chemicals From Water A Hypothesis for the Mechanism
of the Particle Concentration Effect. Chemosphere 16:745-
757.
Mackay, D. M., D.L. Freyberg, P.V. Roberts, and J.A. Cherry.
1986. A Natural Gradient Experiment on Solute Trans-
port in a Sand Aquifer, 1. Approach and Overview of
Plume Movement. Water Resources Research
22(13):2017-2030.
McKay, L.D. and M.R. Trudel. 1987. Sorption of Trichloro-
ethylene in Clayey Soils at the Tricil Waste Disposal Site
near Sarnia. Ontario. Unpublished report. University of
Waterloo Institute for Ground Water Research.
167
-------
Myrand, D., 1987. Diffusion of Volatile Organic Compounds
in Natural Clay Deposits. M. SC. Thesis, Department of
Earth Sciences, University of Waterloo, Waterloo, Ontario.
Nelson, D.W. and L.E. Sommers. 1982. Total Carbon, Or-
ganic Carbon, and Organic Matter. In: Methods of Soil
Analysis, Part 2, Chemical and Biological Properties,
A.L. Page, R.H. Miller, and D.R. Keeney (eds.), ASA
Monograph No. 9, American Society of Agronomy, Madi-
son, WI, pp. 539-580.
Nkedi-Kizza, P., P.S.C. Rao, and A.G. Hornsby. 1985. Influ-
ence of Organic Cosolvents on Sorption of Hydrophobic
Organic Chemicals by Soils. Environ. Sci. Technol.
19:975-979.
Palmer, C.D. and R.L. Johnson. 1989a. Physical Processes
Controlling the Transport of Contaminants in the Aque-
ous Phase. In: Transport and Fate of Contaminants in the
Subsurface, EPA/625/4-89/019, pp. 5-22.
Palmer, C.D and R.L. Johnson. 1989b. Physical Processes
Controlling the Transport of Non-Aqueous Phase Liquids
in the Subsurface. In: Transport and Fate of Contamin-
ants in the Subsurface, EPA/625/4-89/019, pp. 23-27.
Palmer, C.D. and R.L. Johnson. 1989c. Determination of
Physical Transport Parameters. In: Transport and Fate of
Contaminants in the Subsurface, EPA/625/4-89/019, pp.
29-40.
Pankow, J. 1984. Groundwater Contamination by Organic
Compounds: Principles of Contaminant Migration and
Determimtion. Short Course Notes, Oregon Graduate
Institute, Beaverton, Oregon.
Powell, R.M., B.E. Bledsoe, G.P. Curtis, and R.L. Johnson,
1989. Interlaboratory Methods Comparison for the Total
Organic Carbon Analysis of Aquifer Materials. Environ.
Sci. Technol. 23(10): 1246-1249.
Rae, P.S.C., A.G. Hornsby, D.P. Kilcrease, and P. Nkedi-
Kizza. 1985. Sorption and Transport of Toxic Organic
Substances in Aqueous and Mixed Solvent Systems. J.
Environ. Quality 14:376-383.
Roberts, P. V., M.N. Goltz, and D.M. Mackay. 1986. A Natu-
ral Gradient Experiment on Solute Transport in a Sand
Aquifer 3. Retardation Estimates and Mass Balances for
Organic Solutes. Water Resources Research 22(13):2047-
2058.
Schellenberg, K.C., C. Leuenberger, and R.P. Schwarzenbach.
1984. Sorption of Chlorinated Phenols by Natural Sedi-
ments and Aquifer Materials. Environ. Sci. Technol.
18:1360-1367.
Schwarzenbach, R., and J. Westall. 1981. Transport of Non-
polar Organic Compounds from Surface Water to Ground
Water Laboratory Sorption Studies. Environ. Sci. Technol.
15:1360-1367.
Schwille, F. 1988. Dense Chlorinated Solvents in Porous and
Fractured Media: Model Experiments. Lewis Publishers,
Chelsea, MI.
Sudicky, E.A. 1986. A Natural Gradient Experiment in a Sand
Aquifer Spatial Variability of Hydraulic Conductivity
and Its Role in the Dispersion Process. Water Resource
Research 22(13):2069-2082.
Witkowski, P.J., P.R. Jaffe, and R.A. Ferrara. 1988. Sorption
and Resorption Dynamics of Aroclor 1242 to Natural
Sediment. J. Contaminant Hydrology 2:249-269.
Woodbum, K.B., 1986. Solvaphobic Approach for Predicting
Sorption of Hydrophobic Organic Chemicals on Syn-
thetic Sorbents and Soils. J. Contaminant Hydrology
1:227-241.
Wu, S.-C. and P.M. Gschwend. 1986. Sorption Kinetics of
Hydrophobic Organic Compounds to Natural Sediments
and Soils. Environ. Sci. Technol. 20:717-725.
Zachara, J.M., C.C. Ainsworth, L.J. Felice, and C.T. Resch.
1986. Quinoline Sorption to Subsurface Materials: Role
of pH and Retention of the Organic Cation. Environ. Sci.
Technol. 20:620-627.
Zachara, J.M., C.C. Ainsworth, C.E. Cowan, and B.L. Tho-
mas. 1987. Sorption of Binary Mixtures of Aromatic
Nitrogen Heterocyclic Compounds on Subsurface Mate-
rials. Environ. Sci. Technol. 21:397-402.
Zachara, J.M., C.C. Ainsworth, R.L. Schmidt, and C.T. Resch.
1988. Influence of Cosolvents on Quinoline Sorption by
Subsurface Materials and Clays. J. Contaminant Hydrol-
ogy 2:343-364.
168
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Chapter 11
Physiochemical Processes: Volatilization and Gas-Phase Transport
Carl D. Palmer and Richard L. Johnson
Many nonaqueous phase liquids (NAPLs) are volatile
organic compounds of environmental concern (e.g., chlori-
nated solvents, gasoline). They frequently enter ground-water
systems after they have been spilled on the surface and pass
through the unsaturated zone (Figure 11-1). As these NAPLs
flow through the unsaturated zone, a portion of the liquid
remains behind in fingers at residual saturation, in pools of
material on small heterogeneities, or above the capillary fringe
(e.g., Palmer and Johnson, 1989a Feenstra and Cherry, 1987;
Schwille, 1967 and 1988). The NAPL that remains in the
unsaturated zone is an important source of contamination
because it is dissolved by (1) the passing recharge water, and
(2) the passing ground water as the water table rises. Such
sources of contamination can last for many years and con-
taminate large volumes of ground water. However, in addition
to these pathways, contaminants also can be transported through
the unsaturated zone in the gas phase. This transport pathway
may spread the contaminants over a much broader area of the
aquifer. Another complicating factor is the mass transfer of
the contaminants across the atmosphere-soil boundary. Of
greater interest is the implication that these sources of ground-
water contamination can be quickly remediated by actively
pumping the soil gas and removing the volatile organic con-
taminants from the unsaturated zone to the surface where they
may be treated. As with the transport of dissolved contami-
nants, the design of optimal remediation schemes requires
DNAPL Source
mm
rxxxxxxxxxxxxxxxx
Groundwater
Flow I
Lower
Permeability
Strata
Figure 11-1. Transport of a DNAPL into the subsurface illustrating the distribution of the DNAPL, the dense vapors, and the
dissolved chemical plume (after Feenstra and Cherry, 1988).
169
-------
knowledge of the physiochemical processes that control
transport pathways. Mercer and Cohen (1990) provide a re-
cent review of the literature on properties, models, character-
ization, and remediation of NAPLs.
11.1 Volatilization
Near a NAPL spill, a four-phase system exists. The
phases include (1) the aquifer matrix, (2) the residual soil
water, (3) the NAPL, and (4) the air-filled pore space (Figure
11-2). A volatile NAPL partitions from the NAPL phase into
the gas phase where it then can be transported to other
portions of the unsaturated zone. The vapor pressure of a
particular contaminant, Pt, in the gas phase can be calculated
from Raoult's Law:
(nlV)t = PJ(RT)
[11-2]
Pk-
[11-1]
where Xkis the mole fraction of component k in the NAPL
and P°tisthe vapor pressure above the pure component. For
example, if the NAPL is gasoline, the partial pressure on
benzene, one of the many components of gasoline, is the mole
fraction of benzene in the gasoline times the ideal vapor
pressure above pure benzene. The concentration of the gas in
the soil atmosphere then can be calculated from the ideal gas
law:
where n is the number of mole of component k, V is the
volume of gas, T is the kelvin temperature, and R is the gas
constant (0.082057 liter-arm mole'ddg1).
11.2 Gas-Phase Transport
The movement of the contaminants in the gas phase of
the unsaturated zone can be described by performing a mass
balance on a volume of aquifer (Figure 11-3) in a manner
similar to the approach taken in Chapter 10 (see Section 10.1).
11.2.1 Diffusion
Under nonpumping conditions, Fickian diffusion is the
prime process for gas-phase transport. The mass balance or
transport equation can then be written in its one-dimensional
form as:
a*2
dt
[11-3]
where G is the concentration of the contaminant in the gas
phasa, Oj, is the air-filled porosity, D is the free air diffusion
coefficient and i;is the air-phase tortuosity factor. The T
Figure 11-2. Four-phase system consisting of soil matrix, water, NAPL and air (after Schwille, 1988).
170
-------
Gas Phase Transport
Flux In
Diffusive
Flux
Flux Out
Diffusive
Flux
Mass Balance Equation
Change in
Mass Per
Unit Time
Diffusion
Term
Figure 11-3. Mass balance equation for the transport of contaminants in the gas phase.
term accounts for the diffusion taking place in a porous
medium rather than in an open air space such as a room. The
individual molecules must travel around the sand grains and
water films that are present in the porous medium. The term
on the right-hand side of eq. 11-3 represents the net diffusive
flux per unit time of the contaminant in and out of the volume
of soil. The term on the left-hand side of eq. 11-3 represents
the change in mass of gas within the volume per unit time.
Again, this equation is a useful example, listing the minimum
parameters that must be obtained if vapor transport is to be
described.
1
0.8
0.6
0.4
0.2
The air-filled porosity is a physical parameter that can be
obtained by finding the difference of the porosity and the
volumetric water content by using the methods referred to in
Chapter 5 and in Palmer and Johnson (1989b). The air tortuosity
factor can be obtained from empirical equations that are Figure 11-4.
provided from detailed studies of gas-phase transport. For
example, one such equation is the Millington-Quirk
(Millington, 1959) equation:
0.1 0.2
0.3
0*
0.4
0.5 0.6
The air tortuosity factor based on the Millington-
Quirk equation (Millington, 1959) as a function of
air porosity for four different total porosities.
2.333 2
Ta = da I Bt
where 6 is the total porosity,, which is equal to the sum of the
air-filled porosity and the volumetric water content (0a+9w).
So the air-phase tortuosity can be calculated from physical
parameters that are already obtained. The actual value of the
air-phase tortuosity factor varies from 0 when the entire pore
space is occupied by water (saturated conditions) to about 0.8
when the porosity is high and the medium is dry (Figure 11 -
4).
The third important parameter, the free air diffusion
coefficient, sometimes can be found for the specific com-
pound of interest in reference tables. If the diffusion coeffi-
cient for the specific compound (D,) cannot be found, it can
be estimated from the diffusion coefficient (D2) and molecular
weight (M2) of another compound and the molecular weight
of the compound of interest (M,) by:
11.2.2 Gas-Phase Retardation
The example given in eq. 11-3 describes the diffusive
transport of a gas-phase contaminant moving through the air-
filled pores of the unsaturated zone, but it does not account for
any chemical interactions with either the soil water or the soil
matrix. The expression should be modified by adding a gen-
eral reaction term, RXN, to account for these processes. At
equilibrium, partitioning of the contaminant between the gas
171
-------
phase and the water phase is defined by the dimensionless
Henry's constant, KH:
= KaC
[11-6]
where G is the gas concentration and C is the water-phase
concentration. For most neutral, nonpolar hydrophobic or-
ganic compounds, the partitioning between the water phase
and the soil organic carbon can be described by a linear
isotherm (S = KPC) that can be written in terms of the gas-
phase concentrations as:
S = KGIKU
[11-7]
where S is the concentration of contaminant on the soil and Kp
is the linear partition coefficient from the water phase to the
soil organic carbon. Including the reaction term, RXN,
—
dt
or
= QW^L ^L+ pb^L lil^l [n_8b]
3G dt dC dG dt
and recognizing that
3G
coefficient, Kp, are discussed in Chapter 10. The remaining
parameter, KH, can be obtained from tables of chemical prop-
erties (Mabey et al, 1982).
A more physical interpretation to this air-phase retarda-
tion factor, R,:can be given. As the contaminated gas diffuses
through the air-filled pores, the rate of diffusion of the con-
taminant in the air phase is less than that of the air itself
because of the loss of mass from the air phase. This mass is
lost from the residual water contained within the pore space
and/or the soil organic carbon that is part of the soil matrix.
The retardation factor, therefore, can be defined as the ratio of
the rate of diffusion of the air to the rate of diffusion of the
contaminant front in the soil atmosphere. R,. is also the mini-
mum number of pore volumes that must pass through a three-
phase contaminated soil system (soil, water, and air) to remove
the contaminants. It is a minimum because the approach
ignores the effects of mass transfer limitations between phases.
The effects result from heterogeneity and kinetics and unequal
travel times along flow lines from the edge of the contami-
nated area to the vapor extraction well.
Another, more direct, method for obtainingc^D^ (an
effective diffusion coefficient) is through column tests such as
those used by Johnson et al. (1987). These column tests
(Figure 11-5) use a dead-end column with a mixture of
nitrogen and the organic contaminants maintained at one end.
The only process that can carry the contaminant into the
column is molecular diffusion. If a sampling line is fitted to
the interior of the column, then samples can be obtained over
time and the concentration breakthrough curve obtained. This
curve can be fitted to a one-dimensional analytical solution to
the diffusion equation to obtain a fitted, effective diffusion
equation.
as
the addition of these two sinks (sources) of contaminant
within the volume of aquifer (Figure 11-2) results in a mass
balance equation (paradigm):
a2G
ar
[11-10]
where Rjis a gas-phase retardation factor that is defined by
[11-11]
where rb is the dry bulk density of the soil. This retardation
factor is a constant if the water content of the soil does not
change and is analogous to the retardation factor, R, for the
movement of organic contaminants in the saturated zone. The
second term ineq. 11-11 represents the partitioning of the
contaminant from the gas phase to the water phase. The third
term represents the partitioning from the gas phase, through
the water phase, to the solid phase. In this modified example,
the retardation factor, R,, must be determined. Methods for
determining the physical parameters 0m, 9v, and pb already
have been identified. Methods for obtaining the partition
11.2.3 Processes Affecting Gas-Phase Transport
Some insight into the migration of contaminants in the
vapor phase can be attained by considering the different
processes included in eq. 11-9. If there is no partitioning of the
contaminant between the gas phase and the soil (Kpis zero),
and if the Henry's constant, KH, is large (i.e., there is no
Sample Line to GC
Nitrogen + Organics In
Nitrogen + Organics Gut
Figure 11-5. Column for measuring effective vapor phase
diffusion coefficients.
172
-------
significant mass loss to the water phase), then the retardation
factor is close to unity and the contaminants move through the
porous medium with the air. As the value of KHbecomes
smaller, the retardation factor at a given porosity and water
content becomes larger because of the partitioning into the
water phase (Figure 11-6). For example, the amount of retar-
dation for benzene increases with increasing water content
(Figure 11-7) because it partitions into the water phase (i.e.,
KHis small). In contrast, the retardation factor for pentane is
insensitive to changes in volumeric water content because it
does not significantly partition into the water phase (Figure
11-7).
These effects are also seen in column tests (Johnson et al.
1987). The breakthrough curves for methane, trichloroethene
(TCE), and chlorobenzene were obtained in a sand-filled
column under both dry and wet conditions (Figure 11-8).
Methane, with the largest Henry's constant and smallest Koc
values of the three compounds is observed to be the frost to
break through. TCE and chlorobenzene break through later
because of the larger Kocand smaller KHvalues. The differ-
ence between the damp sand and the dry sand reflects the
differences in the Henry's constants for the compounds. In
another test, two columns were prepared, one containing
virtually no soil organic carbon (SOC) and another containing
approximately 1 percent SOC. The breakthrough curves for
methane, octane, and benzene in these two columns (Figures
11-9 and 11-10) demonstrate the role of SOC in the retarda-
tion of the compounds. The differences in the breakthrough
curves for the three compounds in the column containing no
SOC (Figure 11 -9) can be attributed to the differences in the
Henry's constants. The column containing 1 percent SOC
requires more pore volumes to achieve breakthrough of the
octane and benzene, and the differences between the com-
pounds are much greater (Figure 11-10). This increased retar-
dation is the result of the greater partitioning of the contaminant
from the gas phase to the SOC with the larger Koc values for
these compounds.
Temperature can have a significant influence on the rate
of migration of gas-phase volatile organic contaminants. The
diffusion coefficients increase with increasing temperature.
The effect of this temperature dependence can be calculated
from:
D,/D2 = (T/TJT
[11-12]
where T is the kelvin temperature. The exponent, m, should
theoretically be 1.5; however, experimental data yield values
between 1.75 and 2.0 (Hamaker, 1972). The temperature also
affects the vapor pressure of the compounds (Figure 11-11)
and, therefore, the concentration in the gas phase (eq. 11-2).
The Henry's constant also shows a temperature dependence
by increasing with increasing temperature (Figure 11-12).
From the definition of the gas-phase retardation factor (eq. 11-
11), the increased Henry's constant is reflected as a decrease
in R,(Figure 11-13). Thus, fewer pore volumes of air need to
be moved through a contaminated soil to remove the contami-
nant at 35°C than at 10°C.
The concentration of volatile organic contaminants in the
gas phase of the unsaturated zone is influenced by the pres-
40
30
i-
10
Hexane
Figure 11-6.
0.2 0.4 0.6 0.8 1
Henry's Constant (dimensionless)
Retardation factor as a function of dimensionless
Henry's constant when there is no adsorption
(after Johnson et al., 1987).
14
12
10
a
Total Porosity = 0.35
Benzene
Pentane
0.05
0.1 0.15
Water Content
0.2
0.25
Figure 11-7. Retardation factor versus water content for
benzene and pentane (after Johnson et al., 1987).
173
-------
7000
2000
Time (Mm)
3000
4000
Figure 11-8. Relative gas phaae concentration versus time in a column experiment (after Johnson et al., 1987).
0.8
0.6
0.4
0.2
10% Water Content
Octane
Benzene
Methane
6 8 10 12
Pore Volumes
14
16
Figure 11-9. Relative gas phase concentration versus the number of pore volumes of air moved through an unsaturated column
with no soil organic carbon (after Johnson et al., 1987).
ence of boundaries that can impede the rate of migration.
Spills of solvents and hydrocarbons often occur in industrial
and urban areas where parking lots, roads, and foundations
can act as low permeability boundaries that limit the mass
transfer of the contaminated gases from the unsaturated zone
to the atmosphere. In the absence of such barriers, concentra-
tions of the contaminants in the soil-gas phase should remain
low very near the surface. When these barriers are present, the
concentrations in the soil-gas phase can be much greater. The
effect of these impermeable caps is illustrated in the numeri-
cal simulations by Baehr (1987). The mass of total hydrocar-
bon that is in the soil-gas phase is about 2.5 times greater
when there is a cap present than when there is no cap (Figure
11-14).
Measuring permanent gases such as O2and CO2in addi-
tion to the priority pollutants can provide insight into pro-
cesses that are occurring in the subsurface. Measurements of
the distribution of total hydrocarbons in the unsaturated zone
near an oil spill in Bemidji, Minnesota, (Hult and Grabbe,
1985) show that the concentrations are greatest near the
source and decrease with greater distance from the pooled
material (Figure ll-15a). O2is near atmospheric values far
from the source and nearly depleted near the spill area (Figure
ll-15b). CO2distributions are opposite to those of oxygen,
with the highest concentrations being near the source (Figure
ll-15c). While the diffusion and retardation processes dis-
cussed above have an important role in the distribution of the
total hydrocarbons, the depletion of 02and the generation of
174
-------
10% Water Content, 1%SOC
0.3
0.2
Henry's Constant for TCE as a
Function of Temperature
12 16 20 24 28
Temperature (°C)
32 36
Figure 11-10. Relative gas phase concentration versus the
number of pore volumes of air moved through an
unsaturated column with 1% soil organic carbon
(SOC) (after Johnson et al., 1987).
100
•^ 80
I-
1 40
Q.
I" 20
0
TCE Vapor Pressure as a
Function of Temperature
8 12 16 20 24
Temoerature f°C]
28
Figure 11-11. The vapor pressure of TCE as a function of
temperature.
CO2near the source area suggest that biodegradation of the
hydrocarbons is also Occurring (see Chapter 13).
There are other factors that affect the migration of or-
ganic vapors. Cultural factors such as underground utility
conduits, trenches, sewers, and pipes can act as preferential
pathways along which these gases may travel. The type of
backfill used around underground storage tanks affects the
water content and retardation of the gas-phase contaminants.
Other environmental factors such as variations in atmospheric
pressure, fluctuations in the elevation of the water table, and
the amount of infiltration in the contaminated area also have a
significant influence on transport of contaminants in soil gas
at certain sites.
11.3 Vapor Extraction
Vapor phase extraction is an important method for re-
moving residual volatile organic solvents from the subsurface.
Figure 11-12. Dimensionless Henry's constant for TCE versus
temperature.
3.5
a 3
i2 2.5
•I 2
a
1.5
1
0.5
Vapor Retardation Factor for TCE as a
Function of Temperature
12 16 20 24 28
Temperature fC)
32 36
Figure 11-13. Vapor retardation factor for TCE versus tempera-
ture.
In principle, the technique works by removing the volume of
contaminated air from the subsurface. As more air moves into
the contaminated area, the contaminants partition from the
NAPL to the air phase. The extraction is continued until
sufficient pore volumes of air have passed through the con-
taminated zone to remove the entire mass of the NAPL from
the subsurface.
Many of the same problems that are encountered in
ground-water pump-and-treat systems also are expected in
vapor extraction systems. As the contaminated air is extracted
from the unsaturated zone, the highly contaminated soil gas
that was initially present is removed. The concentrations may
then begin to decrease and remain at some concentration that
is substantially lower than the initial concentration but signifi-
cantly higher than the target level (Figure 11-16). This "tail-
ing" is the result of several processes. One factor is the rate of
resorption of the organic contaminants from the soil organic
matter (Figure 11-17). Although little is known about these
rates, some data on hexachlorobenzene (Karickhoff and Morris,
1985) show that the rate is initially rapid and decreases with
175
-------
Total Hydrocarbons
1000
Time (Days)
2000
Figure 11-14. Comparison of total hydrocarbons present in a
soil with a cap versus without a cap (after Baehr,
1987).
time, and the equilibration may take more than 30 days (e.g.,
Figure 10-11, Chapter 10); this time scale is significant com-
pared to the rate of movement of the air. Another consider-
ation is the form of the isotherm itself. If the isotherm is
nonlinear (Langmuir- or Freundlich-type isotherm with the
exponent less than unity), there can be tailing in the concen-
tration versus time curves.
If the NAPL in the porous media is locally surrounded by
water, then the concentrations in the air that is being advected
through the adjacent pores may be limited by the rate of
diffusion through the water (Figure 11-18). In the air, velocity
is low relative to the rate of diffusion (this is the case when
there is no extraction), and the concentrations are limited by
the vapor pressure of the compound. If the air velocities are
large relative to the rate of diffusion, then the concentration of
the contaminant is limited by diffusion through the water. An
analogous situation may arise when thick pools of NAPL are
being removed by vapor extraction. The more volatile compo-
nents from the upper surfaces of the NAPL are removed first.
If the air velocity is large relative to the rate of diffusion of
those volatile components through the NAPL, then the con-
centrations in the gas phase are limited by this rate of diffu-
sion through the NAPL.
Another important aspect of NAPLs that are composed of
more than one component is that as the more volatile compo-
nents are removed, their concentration in the NAPL (as a mole
fraction) decreases. This decrease in the mole fraction de-
creases the vapor pressure (eq. 11-2) and hence the gas-phase
concentration.
Soil heterogeneity plays a major role in controlling the
concentration of contaminants in gases extracted from the
unsaturated zone. As NAPLs infiltrate into the subsurface
they spread into pools on top of lenses of finer grained
material within the aquifer. The NAPL also may be drawn
into the fine-grained zones by capillary action: As air is
advected through the contaminated soil (Figure ll-19a), those
parcels that pass through the fingers or very close to the pools
are close to saturation with respect to the NAPLs. The concen-
tration of the volatiles in the parcels of air between the fingers
and pools is controlled by the rate of vapor diffusion from the
Total Volatile Hydrocarbons (g/m3)
B
1420
1410
1400
1390
1380
Oxygen (ATM)
0.20 —
0.21 —.
Carbon Dioxide (ATM)
1
8
(0
§
o
4
^
I
1420 F
1410
1400 -
1390 •
1380
0.03
0.02
0.01
0.005
Figure 11-15. Distribution of gases near an oil spill in
Bemidjl, MN (after Hult and Grabbe, 1985).
176
-------
On
Off
Max
Si
•c 2
IS
fs
"Residual"
Concentration
\
Residual
Water
Advection
'7
- Time-
Tailing in the vapor concentration versus time
curve for a vapor extraction well.
\
Molecular
Diffusion
Advection
Advection
Organic Carbon or
Mineral Oxide Coating
Equilibrium Concentration
Slow
Desorption
Initial
Rapid
Desorption
Time
I
fi
I
Vapor
Pressure
Limited
Diffusion
Limited
Air Velocity
Figure 11-18. The gas phase concentration controlled by
either the vapor pressure or the rate of diffusion
of the volatile organic through the water phase.
P denotes the residual product.
Figure 11-17. The concentration in the gas phase controlled by
the rate of desorption.
pools and fingers. During vapor extraction, the fingers are
likely to be removed before the pools (Figure ll-19b), be-
cause there is generally less mass of NAPL in the fingers than
in the pools. Also, a greater surface area of the NAPL exposed
in the fingers facilitates the mass transfer to the advected air.
As the residual fingers are removed, the concentrations in the
extracted air should decrease as the less contaminated air
between the pools is mixed with the more contaminated air
from just above the pools. Removal from the pools may
require a substantial period of time (years) to complete. The
NAPL that moved into the finer grained zones is effectively
removed from the advective air flow. Under these conditions,
the removal of NAPL is limited by the relatively slow diffu-
177
-------
B
Source
Figure 11-19. The transport of air through a heterogeneous media with fingers and pools of NAPL present (A) and at a later time
when only pools are present (B).
sion of the NAPL out of the finer grained zone. If a substantial
mass of NAPL is trapped in this way, remediation could
require many years to complete.
11.4 References
Baehr, A.L. 1987. Selective Transport of Hydrocarbons in the
Unsaturated Zone Due to Aqueous and Vapor Phase
Partitioning. Water Resources Research 23(10): 1926-
1938.
Feenstra, S. and J.A. Cherry. 1987. Dense Organic Solvents in
Ground Water An Introduction. In: Dense Chlorinated
Solvents in Ground Water, Progress Report 0863985,
Institute for Ground Water Research, University of Wa-
terloo, Waterloo, Ontario.
Hamaker, J.W. 1972. Diffusion and Volatilization. In: Or-
ganic Chemicals in the Soil Environment Vol. 1., C.A.I.
Goring and J.W. Hamaker (eds.), Marcel Dekker, New
York, Chapter 5.
Huh, M.F. and R.R. Grabbe. 1985. Distribution of Gases and
Hydrocarbon Vapors in the Unsaturated Zone. In: Pro-
ceedings, U.S. Geological Survey Second Toxic Waste
Technical Meeting, Cape Cod, MA, October 1985, U.S.
Geological Survey Open File Report 86-0481, pp. C21-
C25.
Johnson, R.A., C.D. Palmer, and J.F. Keely. 1987. Mass
Transfer of Organics Between Soil, Water, and Vapor
Phases: Implications for Monitoring, Biodegradation and
Remediation. In Proc. NWWA/API Symp. on Petroleum
Hydrocarbons and Organic Chemicals in Ground Wa-
ter-prevention, Detection and Restoration, National
Water Well Association, Dublin, OH, pp. 493-507.
Karickhoff, S.W. and K.R. Morris. 1985. Sorption Dynamics
of Hydrophobic Pollutants in Sediment Suspensions.
Environ. Toxicol. Chem. 4:469-479.
Mabey, W.R., et al. 1982. Aquatic Fate Process Data for
Organic Priority Pollutants. EPA 440/4-81/014 (NTIS
PB87-169090), Chapter 4.
Mereer, J.W. and R.M. Cohen. 1990. A Review of Immiscible
Fluids in the Subsurface Properties, Models, Character-
ization and Remediation. J. Contaminant Hydrology 6:107-
163.
Millington, R.J. 1959. Gas Diffusion in Porous Media. Sci-
ence 130100-102.
Palmer, C.D. and R.L. Johnson. 1989a. Physical Processes
Controlling the Transport of Non-Aqueous Phase Liquids
in the Subsurface. In: Transport and Fate of Contami-
nants in the Subsurface, EPA/625/4-89/019, pp. 23-27.
Palmer, C.D. and R.L. Johnson. 1989b. Determination of
Physical Transport Parameters. In: Transport and Fate of
Contaminants in the Subsurface, EPA/625/4-89/019, pp.
29-40.
Schwille, F. 1967. Petroleum Contamination of the Subsoil -
A Hydrological Problem. In: The Joint Problems of the
Oil and Water Industries, Peter Hepple (ed.), Elsevier,
New York, pp. 23-53.
Schwille, F. 1988. Dense Chlorinated Solvents in Porous and
Fractured Media. Lewis Publishers, Chelsea, ML
178
-------
Chapter 12
Physiochemical Processes: Inorganic Contaminants
Carl D. Palmer and William Fish
Inorganic compounds are a common and widespread
class of contaminants at many hazardous waste sites. The
relative importance of these inorganic contaminants is illus-
trated in the Records of Decision (RODS) signed by EPA
between 1982 and 1986 (Booz-Allen and Hamilton, Inc.,
1987). Of the 108 RODS, 56 percent involved Superfund sites
where inorganic compounds were designated as a potential
threat or problem (Palmer et al., 1988). While organics/
volatile organic compounds (VOCs) are the most frequently
reported contaminants, heavy metals and inorganic are the
second and third most frequently reported categories of haz-
ardous substances (Figure 12-1). Inorganic waste problems
Primary Hazardous Substances Detected
Acids
Arsenic
Asbestos
Carcinogenic
Chromium
Dioxin
Heavy Metals
Inorganics
Mining wastes
Oils
Organics/VOCs
PAHs
PCBs
PCEs
Pesticides
Phenols
Radioactive
Sludge
Solvents
Synfuels
TCE
Toulene
(
• i • i
\2
36
x i ?y
160
35
" • 'j 126
"• \20
"\42
; — ~\23
' \24
\ \23
35
1
.:::::J S2
—J3\
) 50 100 15t
can be particularly severe because they often occur at sites
that cover many square miles. Remediation of such sites is
often difficult simply because of the size of the affected area.
The most common inorganic constituents of concern are
the 13 priority pollutant metals (Table 12-1). However, other
inorganic substances such as nitrate, phosphate, cyanide, and
radionuclides may be found at levels far exceeding drinking
water standards. Iron, manganese, aluminum, calcium, silica,
and carbonates are not priority pollutants but also can contrib-
ute to the overall cost of a remediation scheme by increasing
maintenance and disposal costs. For example, air strippers at
several sites have been temporarily disabled by iron precipita-
tion problems.
The behavior and toxicity of inorganic compounds are
affected by chemical and physical processes. Understanding
these processes may lead to more cost-effective restoration of
contaminated sites and can reveal how inorganic substances
may affect the cleanup of sites, even at sites where organic
contaminants are the main concern.
Table 12-1. The 13 Priority Metals
Figure 12-1. Primary hazardous substances detected at
hazardous waste sites. Based on data from Booz-
Allen and Hamilton, Inc. (1987), from Palmer et al.
(1988).
Silver
Arsenic
Barium
Cadmium
Chromium
Nickel
Mercury
Lead
Selenium
Thallium
Antimony
Copper
Zinc
179
-------
12.1 Chemical Processes
Several chemical processes affect the concentration, spe-
cific form, rate of transport, and ease of removal of inorganic
substances from the subsurface. These processes include (1)
chemical speciation, (2) oxidation/reduction, (3) dissolution
and precipitation of solid phases, (4) ion exchange and ad-
sorption onto the soil matrix, and (5) transport of particles in
the subsurface. One of the difficulties in working with inor-
ganic contaminants is that all of these processes can be
operating simultaneously. Therefore, it is sometimes difficult
to determine which is the most important at a particular site.
The relative importance of these processes not only varies
from site to site but may vary from one area to another within
a given site. The relative importance of these processes also
may change during the cleanup operation as subsurface chemi-
cal conditions are altered. Each of these processes is discussed
below.
12.1.1 Speciation
When a water sample is sent to the laboratory for metals
analysis, the results are usually returned as total concentra-
tions. In reality, the metals interact with anions (or ligands)
that exist in the ground water to form different chemical
species or "aqueous complexes," For example^ cadmium may
exist in solution as Cd2+, CdCl+, CdCl°, CdOH*, or as other
complexes. In addition to Cl and OH, petal ions. can. com-
bine with SOl; CO*, P, S2', NH i Poj', CN", or polyphos-
phates. Complex formation of transition metals is more
extensive than for other metals. The relative tendency of
metals to form complexes is in the order
Fe(III) > Hg > Cu > Pb > Ni >
Zn > Cd > Fe(II) > Mn > Ca > Mg
(Hanzlik, 1976). The concentration of aqueous complexes
depends upon the concentration of the metal ion of interest,
the ligand with which it forms the complex (Figure 12-2), and
1.0
0.8
3
I
o
0.2
0.0
-2
-1 0
Log 1C!-}
the concentration of the other metal ions that may compete for
the ligand. Because chemical reactions usually are governed
by the amount of free ion rather than by the total metal,
knowledge of the concentration of complexes is important to
properly identify the role of processes such as adsorption,
mineral dissolution, and precipitation.
The formation of inorganic aqueous complexes is often a
rapid process that can be described by equilibrium expres-
sions. For example, the formation of the mercuric chloride
complex, HgCl+, can be written as the chemical reaction:
Hg2++ cr+ Hgcr
which can then be written in terms of an equilibrium constant
as:
K= {Hg2+} {Cl-}/{HgCn
[12-1]
Figure 12-2. Fraction of total Cd in various chemical com-
plexes versus the Cl concentration (after Moore
and Ramamoorthy, 1984).
where the braces represent the thermodynamic concentration
or "activity" of the chemical species. The activity of the
species is simply the product of the concentration of that
chemical species and an "activity coefficient," y. For example,
the activity of Hg2+can be written as:
[12-2]
where K,, is an equilibrium (or stability) constant for the above
reaction and the brackets indicate the concentration of the
chemical species. The activity coefficient accounts for the
change in chemical reactivity due to electrostatic interactions
among ions in solution. Several methods are used to calculate
activity coefficients. See Stumm and Morgan (1981), Morel
(1983), or Sposito (1986) for more detailed discussions of
these methods.
Calculation of the concentrations of each chemical spe-
cies in solution requires the total concentration of each metal
and each ligand in solution as well as the equilibrium constant
for the formation of each complex. Total concentrations are
obtained from chemical analysis of the water sample. Equilib-
rium constants (or more fundamental thermodynamic data)
for inorganic complex formation have been studied by many
researchers and are tabulated in a variety of references (e.g.,
Ball et al, 1980; Felmy et al, 1984; Smith and Marten, 1976).
An extensive list of thermodynamic data sources is given by
Nordstrom and Munoz (1985).
While in principle it is easy to calculate the concentration
of these aqueous complexes, it does not take too many combi-
nations of metals and ligands to result in a series of equations
that are unmanageable to solve by hand. Fortunately, compu-
tational algorithms can quickly perform these calculations, so
time can be better invested in interpreting the results of the
calculations.
12.1.2 Dissolution/Precipitation
Ground water passing through an aquifer may be in direct
contact with a wide variety of mineral phases. Dissolution or
weathering of those mineral phases contributes greatly to the
natural chemical composition of the ground water. Generally,
180
-------
dissolution refers to a reaction where all of the chemical
species that comprise the mineral come into solution. Some
common minerals that undergo dissolution in shallow aquifer
systems include gypsum (CaS04« F^O), calcite (CaCO3) and
quartz (SiO?). Weathering is a partial dissolution process in
which certain ions come into solution while others remain as
part of secondary mineral phases. During the initial stages of
weathering of aluminosilicates, Ca2+, Mg2+, Na+, K+, and some
H4SiO go into solution while the remaining ions become clays
such as kaolinite and montmorillonite, If the concentrations of
certain ions are sufficiently high, they may be removed from
solution by the formation of a solid phase (i.e., by precipita-
tion). These precipitated minerals may dissolve later if physi-
ochemical conditions within that portion of the aquifer change.
These types of reactions can have a great effect on the
concentrations of priority pollutants within the aquifer. For
example, the weathering of fly-ash piles can yield selenium,
arsenate, lithium, and heavy metals (Honeyman et al., 1982;
Murarka and Macintosh, 1987). BaCr04may precipitate in an
aquifer contaminated with Cr04and later it may dissolve
during remediation, thus prolonging the time required to
restore the aquifer. Pb can be removed from solution by
precipitation as PbCOr In addition, dissolution of naturally
occurring minerals can neutralize acid or alkaline waters and
thus enhance the adsorption or precipitation of priority metals.
The precipitation of calcite or hydrous ferric oxides may
reduce the permeability of the aquifer, clog well screens, and
increase the cost of treatment and disposal. Therefore, it is
advantageous to be aware of the potential for these types of
reactions and to include their effects in the cost and design of
remediation.
Equilibrium between the ground water and a solid phase
can be expressed in terms of an equilibrium constant (or
volubility product). For example, the dissolution or precipita-
tion of BaCr04, which involves two priority metals, is written
as:
BaCrO4(s) <-» Ba2* + CrO*'
and at equilibrium this can be expressed as:
K,=
[12-3]
where Kspis the volubility product. As with the aqueous
complexes, the terms in the braces refer to the activity of the
particular species and not the total concentration of the ele-
ments. The equilibrium constants for many solid phases can
be found in the same references given for the aqueous com-
plexes. The ability of a ground water to dissolve or precipitate
a solid phase is sometimes expressed as a saturation index
(SI), which is defined as:
SI =
[12-4]
where the IAP or "ion activity product" is the same expression
of ion activities used for the volubility product but at the
concentrations found in the ground water rather than at equi-
librium. For example, the ion activity product for barium
chromate is (Ba2+) {Cro2"4}. If the ground water is in equilib-
rium with the solid phase, then the IAP is equal to the Kspand
the saturation index is equal to zero. If the saturation index is
less than zero, the water is undersaturated with that solid
phase and the ground water has the potential to dissolve that
phase if it is present. If the saturation index is greater than
zero, then the water is supersaturated with respect to the solid
phase and has the potential to precipitate that phase provided
that the reaction is fast enough to occur with the time scales of
interest. Calculation of the saturation indices for mineral
phases requires knowledge of the concentration of the aque-
ous complexes. Therefore, the computational tools used to
calculate the concentration of aqueous complexes also can be
used to calculate saturation indices.
Saturation indices at waste sites can be useful for identi-
fying potential sources and sinks for metal ions. If calculated
saturation indices for PbCO3are close to zero, then Pb is
likely being removed from solution and may not move very
far from the source. If the saturation indices for BaCrO4are
close to zero, there may be a large reserve of solid phase in the
aquifer that could prolong a pump-and-treat remediation
scheme for the removal of the CrO4. If the waters of interest
are undersaturated or supersaturated with respect to solid
phases of interest, the amount of the solid phase that must be
dissolved or precipitated per liter of water to reach equilib
rium can be calculated. Such theoretical calculations may be
particularly useful in evaluating the potential for mineral
precipitation as a result of mixing contaminated and uncon-
taminated waters in extraction wells. Note that saturation
indices alone do not prove the presence or absence of a
mineral phase. However, saturation indices are relatively easy
to obtain and are valuable for identifying possible mineral
phases.
12.1.3 Oxidation/Reduction
The number of electrons associated with an element
dictates its oxidation state. Elements can exist in several
oxidation states. For example, iron commonly exists in the +2
or +3 state, arsenic as +3 or +5, and chromium as +3 or +6.
Oxidation-reduction (redox) reactions involve a transfer of
electrons and, therefore, a change in the oxidation state of
elements. Redox reactions are important to studies of subsur-
face contamination because the chemical properties for the
elements can change substantially with changes in the oxida-
tion state. For example, in slightly acidic to alkaline environ-
ments, Fe(III) is fairly insoluble and precipitates as a solid
phase (hydrous ferric oxide) that has a large adsorption capac-
ity for metal ions. In contrast, Fe(II) is fairly soluble and its
oxides have a much lower adsorption capacity. As the Fe(III)
solid phase is reduced, not only is the Fe(II) brought into
solution but so are any contaminants that may have been
adsorbed onto it (Evans et al., 1983; Sholkovitz, 1985). An-
other environmentally important redox-active element is chro-
mium. Hexavalent chromium, Cr(VI), exists in ground water
as the relatively mobile and toxic anions HCrO4and CrO24.
The reduced form of chromium, Cr(III), is a cation that under
slightly acidic to alkaline conditions is fairly insoluble, readily
adsorbed, and much less toxic than Cr(VI). Selenium also
undergoes important redox transformations. Selenate (Se(VI))
is more mobile and less toxic than selenite (Se(IV)).
181
-------
Because redox reactions involve the transfer of electrons,
the change in oxidation state of one element necessitates a
change in the oxidation state of another. For example, as
Cr(Vl) is reduced to Cr(III), it must gain three electrons from
another element. One possible electron donor is ferrous iron
(Fe(II)):
HCrO + 3Fe2+ + 7H+
3Fe3
4H2O
Redox reactions cannot occur unless there is both a suitable
electron donor and a suitable electron acceptor.
The expected form of an element at equilibrium depends
on the pH and the redox state of the water. The redox state is
measured by an electrical potential (in volts (v) or millivolts
(mV)) at a standard electrode. This potential is called the EH
of the ground water. Alternatively, the redox conditions are
sometimes reported in terms of the "pe," which is the negative
logarithm of the activity of the electron. This is a direct
analogy to pH, which is the negative logarithm of the activity
of H+. The EHand pe of a watar measure the same property,
but due to differences in definition they are not numerically
equivalent. Redox conditions within natural aquifers vary
from highly oxidizing conditions (high EH, - +800-900 mV)
to very reducing conditions (low EH,~ -200 mV). Variation
within contaminated aquifers is at least as great as this range,
but often is marked by abrupt transitions over much smaller
scales than is typical of uncontaminated aquifers (see Sections
8.3.3 and 8.4).
The conditions of pH and EHfor which a particular redox
species is theoretically stable are represented graphically on
an EH-pH diagram. These diagrams are also known as pe-pH
and Pourbaix diagrams. For example, the EH-pH diagram for
Fe (Figure 12-3) illustrates the predominance of the iron
hydroxide solid at slightly acidic to alkaline conditions and
high EHconditions, whereas aqueous Fez+predominates at
low EHand slightly alkaline to acid conditions. Methods for
the construction of such diagrams can be found in Garrels and
Christ (1965), Drever (1989), and Stumm and Morgan (1981).
A collection of EH-pH diagrams for many metals was prepared
byBrookins(1988).
In theory, knowledge of the EH, pH, and total elemental
concentrations in a ground water allows the quantitative pre-
diction of the concentration of each redox-active species in
solution. However, many redox reactions are microbially
catalyzed, nonreversible, and, therefore, not found in a state of
mutual equilibrium. Redox species that are not at equilibrium
with each other often have been observed to occur together
(Lindberg and Runnels, 1984). Except for a very few situa-
tions, it is impossible to predict general redox behavior in
aquifers using equilibrium concepts. Nonetheless, there are
observable and consistent trends in redox conditions in natural
aquifers that suggest that at least qualitative estimates of
behavior are possible. Because of the importance of redox
reactions it is important to know at least the possible changes
in the redox state. Some idea of the transformations that may
be taking place within the aquifer can be obtained in certain
cases (e.g., for Cr, Se, As) by directly measuring the concen-
trations of different oxidation states of the contaminants.
More must be learned about the rates of redox reactions if
24
18
14
12
a
4
8. 0
-4
-8
-12
-16
-20
Fe(OH)2-
0 2 4 6 8 10 12 14
pH
Figure 12-3. pe-pH diagram for the Fe-H,O system.
these reactions are to be put into proper perspective with
regard to the transport and removal of inorganic contaminants
from the subsurface. A review by Fish (1990) summarizes
these concepts, problems, and some possible solutions.
12.1.4 Adsorption/Ion Exchange
Ion exchange and adsorption can exert a great influence
on the concentrations of ions in solution. Clay minerals are
important ion-exchangers in subsurface systems. During ion
exchange, ions in certain layers of the three-dimensional
structure of clays are replaced by ions in solution, while a
constant total charge within the clay layer is maintained. For
example, Ca-Na exchange can be written as:
2NaX + Ca2+ <-> CaX + 2Na+
and using the Vanselow (1932) convention can be expressed
in terms of a "selectivity coefficient", Ks, as:
Ks _
[12-5]
where NaX and CaX represent the Na and Ca on the clay and
XNaXand XCaXare the mole fractions of exchangeable sites
occupied by Na and Ca, respectively. The selectivity coeffi-
cients are empirical and vary with the concentration of the
cations in the ground water (Reichenberg, 1966) so that
location-specific values must be used.
182
-------
Knowledge about ion exchange is important to under-
stand the binding of alkali metals and the alkaline earths and
some anions to clays and condensed humic matter (Sposito,
1984; Helfferich, 1962). However, ion exchange does not
adequately describe the interaction of many transition metals
with mineral surfaces. These interactions are better described
by adsorption processes. Ionic adsorption involves the coordi-
nation bonding of metal cations and anions to surfaces of
minerals exposed in the pore space of the aquifer. In ion
exchange, the total electrostatic charge of the solid phase is
constant, whereas for adsorption, the charge of the surface
varies with solution pH and the amount of ions adsorbed. At
this time, no generally accepted model for the adsorption of
inorganic ions exists, so several approaches are discussed
below.
As with organic solutes, adsorption isotherms for inor-
ganic compounds can be constructed and Langmuir- and
Freundlich-type isotherms can be utilized. In general, anions
tend to follow Langmuir isotherms while cations tend to
follow Freundlich isotherms (Dzombak, 1986). For the or-
ganic contaminants considered in Chapter 10, linear isotherms
were found for many compounds allowing a constant retarda-
tion factor to be applied in transport calculations. There is no
model of comparable simplicity that can properly describe the
transport of inorganic substances in the subsurface. One of the
key reasons why an analogous model does not work for
inorganic constituents is the strong dependence of the amount
of adsorption on the pH of the ground water. This is illustrated
in a "pH-edge" which is a plot of the fraction of the total mass
of a metal adsorbed versus the pH of the solution (Figure 12-
4).
"8
1
100
80
60
40
Cations
+—/ / /-
Increasing Adsorbent
i it
40
Increasing Adsorbent
\ \ \
F i g u B 12-4. pH adsorption edges for cations and anions (after
Dzombak, 1986). The arrows illustrate the shift in
the edge with increased adsorbent concentration.
For cations, very little of the metal is adsorbed onto the
aquifer material at low PH. As the pH increases, the fraction
adsorbed onto the soil increases until virtually all of the metal
is adsorbed or all metal-binding sites are occupied. The exact
position of the pH-edge on the diagram depends on the
specific ion considered, the concentration of the metal, the
amount of adsorbent, and the concentration of other ions in
solution, These additional ions may compete for adsorption
sites or form complexes with the metal ion and shift the edge
(Figure 12-5) (Benjamin and Leckie, 1981). The pH-edges for
anions are approximately mirror images of those for cations,
with maximum adsorption Occurring at low pH and decreas-
ing with increasing pH. Any useful model of ionic adsorption
of metal ions should account for the pH dependence of
adsorption. The so-called "surface complexation models" for
ion adsorption meet this criterion (e.g., Stumm et al., 1976
Schindler, 1981; Schindler and Stumm, 1987; Dzombak and
Morel, 1990). These models have a foundation in chemical
theory and if used for a well-defined system, may be applied
over a somewhat wider range of conditions than the specific
experiments used to determine the particular model param-
eters. Surface complexation models are based on the concept
that ions form complexes with solid-phase atoms at the oxide/
solution interface. These complexes are analogous to com-
plexes formed in solution between metals and ligands. There
is, however, an added difficulty due to the formation of a
surface charge at the oxide-water interface.
An oxide can be viewed as an array of metal ions and
oxygen atoms (Figure 12-6a). When the oxide is immersed in
an aqueous solution, water molecules arrange themselves
around the surface metal ions (Figure 12-6b). Some of these
adsorbed water molecules then dissociate, and the resulting
hydrogen ions bind to the adjacent surface oxygen atoms
(Figure 12-6c). Adsorption of anions and cations to the oxide
surface can be described as an exchange of the metal ions for
the H+and the ligands for the OH groups on the surface of the
•§
2
i
100
80
60
40
20
0.7MNO3
0.5 MCI
0.2 M SO4
0.7M
10
pH
Figure 12-5. The effect of anions on the adsorption of Cd on
SiO, (after Benjamin and Leckie, 1981).
183
-------
H H H H H H H H
H H H H
Figure 12-6. Formation of a hydroxylated oxide surface in
water (after Schindler, 1981).
oxide (Figures 12-7 and 12-8). Such adsorption reactions can
be written in a manner similar to the solution complex reac-
tions discussed above:
+ L- o=L + OH-
where the symbol = denotes the oxide surface, M2+is a
divalent cation, and L is a monovalent anion. Equilibrium
constant expressions for these reactions are
Several surface complexation models have been devel-
oped. The most important feature that distinguishes one sur-
face complexation model from another is the treatment of the
electrostatic term. Computation of the electrostatic effects on
adsorption requires postulation of a particular arrangement of
electrical charges near the surface. These charge distributions
are developed by hypothesizing one or more layers of charges
near the surface. The three types of surface complexation
models often used are two-layer models, Stem-layer models,
and triple-layer models. More detailed discussions of these
models are given by Dzombak (1986) and Dzombak and
Morel (1990). These models represent increasing complexity
in the geometric view of the oxide-water interface and require
an increasing number of fitted parameters in addition to the
equilibrium constants. Despite apparent differences in their
sophistication, all three of these surface complexation models
equally describe acid-base titration data for oxide surfaces
(Westall and Hohl, 1980; Dzombak, 1986). In addition,
Dzombak (1986) found that the simpler two-layer model was
quite capable of modeling anion adsorption and cation adsorp-
tion if the cation concentrations were not extremely high.
Therefore, the choice of model is best based on which is the
most parsimonious; the obvious choice in many cases is the
two-layer model.
One of the major disadvantages to using this type of
adsorption model is a lack of knowledge about the equilib-
rium constants. The problem in using constants reported in the
literature is that they are specific to the adsorption model used
to fit the experimental data. Therefore, reported constants are
quite dissimilar from those appropriate for a different adsorp-
tion model. To overcome this limitation, Dzombak (1986) and
Dzombak and Morel (1990) reinterpreted the raw data from
adsorption experiments for different ions on hydrous ferric
oxides (HFO) using the basic two-layer model. This effort
provides a set of consistent constants based on a common
adsorption model and makes more widespread use of such
models feasible. If the main adsorbent in the aquifer is HFO,
the derived constants provided by Dzombak and Morel (1990)
should be adequate. However, natural porous media may
contain many oxides and other surfaces to which metal ions
a-
M++H
These equations are analogous to the equilibrium expressions
for the aqueous complexes, except for the y., terms which
describe the effect of the electrostatic charge near the surface
of the oxide.
Figure 12-7. Adsorption of a divalent cation on a hydroxylatad
oxide surface (from Palmer et al., 1988).
184
-------
H
Figure 12-8. Adsorption of a monovalent anion on a hydroxy-
latad oxide surface (from Palmer et al., 1988).
can bind (Figure 12-9), and a consistent set of equilibrium
adsorption constants has yet to be derived for other types of
surfaces. It also is unclear whether mixtures of surfaces are
linearly additive.
12.2 Particle Transport
A potentially important mechanism for the migration of
inorganic substances is particle transport. These particles may
be inorganic, organic, or biological and may include bacteria,
viruses, natural organic matter, inorganic precipitates, asbes-
tos, and clay. Inorganic ions may migrate as integral constitu-
ents of the particles or they maybe adsorbed onto the surfaces
of the particles. The distance these particles move depends
upon the size of the particles relative to the size of the pores
through which they must pass as well as the chemical condi-
tions in the subsurface.
Particles can be removed from solution by three major
mechanisms: (1) surface filtration, (2) straining, or (3) physi-
cal-chemical processes (McDowell-Boyer et al., 1986). If the
particles are larger than the largest pores within the aquifer,
they cannot penetrate the aquifer and they are filtered out at
the interface between the medium and the source of the
particles (surface filtration). If the particles are smaller than
the largest pores but larger than the smallest pores, the par-
ticles can travel some distance into the aquifer before they
encounter a pore through which they cannot pass; they will
then be strained from solution (straining). If the particles are
smaller than the smallest pores in the medium, then they can
travel great distances. However, the particles still can be
removed from solution by adhering to the sand grains because
of collision. Collision with the sand grains occurs as a result
of sedimentation, interception, and Brownian motion. Par-
ticles in the subsurface also can aggregate if chemical condi-
tions such as pH or ionic strength change significantly
(physical-chemical processes). The particle aggregates then
can be removed from the water by straining.
Particle transport is not likely to be an important factor in
every environment. Therefore, it is useful to target those
situations where particle transport is most likely to be impor-
tant Such situations include environments where there are
high concentrations of organic carbon, dissolved solids, or
suspended solids. Movement of particles may be induced in
areas where the flow rates are very high, either because of
natural flow conditions or more commonly because of high
pumping rates. Any time there is an abrupt transition in pH or
rcdox conditions within the aquifer subsurface, there is an
opportunity for the precipitation of colloidal-size particles
that can travel through the aquifer. When a water appears to
be supersaturated with common mineral phases that are nor-
mally expected to be at equilibrium, then particle transport
should be suspected. A simple example of the latter would be
high iron concentrations in the presence of oxygen under
mildly acidic to alkaline conditions. The iron is likely to
precipitate as a fine colloid and to be included in the total iron
analysis of the water (Fish, in press). If the particle transport is
believed to be important, there are several techniques for
particle detection. These techniques include filtration, micros-
copy, electrophoresis, and light scattering. A review of light-
scattering techniques is provided by Rees (1987).
Particle transport has been documented for at least two
different ground-water contamination sites. It has been ob-
served in a contaminant plume emanating from rapid infiltra-
tion beds used to recharge treated sewage to a sand and gravel
aquifer at Otis Air Force Base on Cape Cod, Massachusetts
(Gschwend and Reynolds, 1987). Particles near the source
were relatively small (< 6nm) but down-gradient from the
infiltration beds, apparently mobile particles about 100 nm in
diameter were observed (Figure 12-10). Chemical analysis of
the particles indicated that they were composed of Fe(II) and
PO4and may be the mineral vivianite (Fe,(PO4)2). As the
treated sewage, which was high in organic carbon, was re-
charged to the aquifer, reducing conditions were created within
the plume. Reductive dissolution of the naturally existing
HFO or other iron-containing minerals resulted in elevated
levels of Fez+ in solution. Phosphate entered the aquifer in
wastewater percolating through the infiltration basins. The
phosphorus originated from detergents used prior to the mid-
1970s. As phosphate entered the aquifer with the recharge
water and mingled with Fe(II), the water eventually reached
saturation and began to precipitate the Fe-phosphate solid
phase (Figure 12-11). However, the precipitate remained in
solution as fine particles that migrated through the sand and
gravel aquifer.
Another documented example of particle transport in-
volves the migration of radionuclides at the Nevada Test Site
(Buddemeir and Hunt, 1988). Large volumes of water were
passed through a series of ultrafilters to measure the concen-
tration of radionuclides in different size fractions of particles.
The results (Figure 12-12) indicate that some nuclides such as
125Sb are almost entirely in solution (<3 nm), while nuclides
such as 54Mn are transported on relatively large particles. Yet
other nuclides, such as 105Ru, are not associated with any
particular particle size but are evenly distributed over the
different particle sizes.
The definition of what is a molecule in solution and what
is a colloidal particle is arbitrary. For many years the 450-nm
pore size was used as the standard break between what is in
solution and what is a particle. However, this pore size
corresponds to molecular weights in the range of several
hundred thousand atomic-weight units. Materials of this size
are now thought to be more properly defined as colloidal
185
-------
Natural Porous Medium
Mineral Edge with Enhanced Adsor^ '.ion
Hornblende
Quartz
Microbes that
Can Act as
Binding Sites
K-Feldspar
Discontinuous
HFO Coatings
Clay
(Kaolinite)
Reduced Surface
Area Due to
Mineral Contact
Points
Mica
Plagioclase
(Solid Solution)
Organic Matter
Figure 12-9. Natural porous media containing many different types of adsorption surfaces (from Palmer et al., 1988).
Rapid
Infiltration
Beds
Size = 6nm
Size = 104 nm
I
Size = 102 nm
\
Average Groundwater
Flow Direction
I
250
Ashumet
Pond
Meters
Figure 12-10. Particle sizes measured in the subsurface at the
Otis Air Force Base, Cape Cod, MA (modified from
Gschwend and Reynolds, 1987).
particles. Consequent^, smaller pore-diameter filters are in-
creasingly used to distinguish "particles" from "solutes." How-
ever, there is no objective standard of what constitutes the
proper dividing size.
Particle transport is particularly important for sites where
the contaminant is highly toxic and the general expectation is
that the contaminant is not mobile because of its high affinity
for adsorption. However, if the material to which it is ad-
sorbed is fairly mobile, then the sorbate may move rapidly
beyond the site.
For remediation efforts, particle transport can be a benefit
or a liability. If particle transport is significant, it may be
possible to remove a significantly greater mass of contamin-
ant per unit time (hence per unit cost) than if the contaminant
were adsorbed onto immobile particles. However, particles
can plug injection wells or aggregate in the subsurface and
reduce the permeability of the formation near the extraction
wells. These effects may increase the overall cost of aquifer
restoration if filter presses and longer pumping schedules are
required to overcome these problems.
12.3 Organic-Inorganic Interactions
Mixtures of many types of wastes are found in landfills,
dumps, and ground-water contamination sites. Consequently,
it is not unusual to find both inorganic and organic contami-
nants together. For example, an analysis of leachate from a
municipal landfill in Brookhaven Town, Long Island, New
York, revealed 660 ug/L Cr (VI), 127 ug/L lead, 151 ug/L
186
-------
POj Enters Subsurface
with Waste Water
Conditions
Ptevaitwithin Plume.
Reduction of
Fe(OH)3 toFe(ll)
3Fe 2++ 2POf — Fe3 (PO4 )2
(vivianite)
Figure 12-11. Formation of particles at the Otis Air Force Base site, Cape Cod, MA.
xylenes, 40 ug/L methylene chloride, 27 ug/L naphthalene,
and 25 ug/L benzene (Black and Heil, 1982). At Woburn,
Massachusetts, Cr (VI) levels of over 2,000 ug/L were mea-
sured in ground-water samples that contained high levels of p-
chloro-m-cresol, phenol, p-nitrophenol, N-nitroso-
diphenylamine, phthalate esters, and 35 other organic com-
pounds (Cook and DiNitto, 1982). The behavior of inorganic
constituents in such waste mixtures can be dramatically dif-
ferent from their behavior when the inorganic contaminants
are found by themselves.
The interactions among organic and inorganic compounds
can be classified as either direct or indirect. Direct interactions
include processes such as complexation (chelation) of metal
ions with organic solutes, oxidation-reduction reactions be-
tween organic and inorganic constituents, and the competition
between organic and inorganic solutes for adsorption sites.
Indirect interactions refer to the changes in pH and redox
conditions as a consequence of degradation of organic con-
taminants in the subsurface. Most of the research on direct
interactions between organic and inorganic materials has fo-
cused on finding or characterizing synthetic pathways for the
commercial production of chemicals, for example, the oxida-
tion of alcohols with Cr(VI) to produce aldehydes or carboxy-
lic acids. Often this research has been conducted under extreme
conditions of concentrations and pH that are of little environ-
mental significance. The few studies that are of environmental
interest indicate that organic-inorganic reactions are impor-
tant in several situations. Stone (1986) found that phenols can
be oxidized in the presence of MnOz. Voudrias and Reinhard
(1986) reviewed several investigations of the oxidation of
organic compounds by metal-substituted clays. Laha and Luthy
(1990) studied the oxidation of aniline and other aromatic
amines by MnOz. Fish and Elovitz (1990) observed the reduc-
tion of hexavalent chrome by cresols. They found that the rate
of reduction was strongly dependent on the pH and the
particular isomer involved in the reaction. While the implica-
tions of these results for remediation have yet to be seriously
considered, the results may have implications on the design of
systems where waters may potentially mix in extraction wells
and treatment trains.
Even at sites contaminated only by organic compounds,
inorganic constituents cannot be ignored. While elevated con-
centrations of inorganic ions may result directly from waste
leachate, they also may result from mobilization of naturally
occurring ions in response to the changing pH and redox
conditions induced by organic contamimnts. Such changing
conditions are typical consequences of biodegradation (Figure
12-13). Biodegradation consumes oxygen, thereby decreasing
the EH(pe) within the contaminant plume. COZ, a by-product
of biodegradation, forms carbonic acid and decreases the pH
within the plume. Organic acid by-products also may decrease
the pH. These processes can result in the resorption of metal
ions and the dissolution of hydrous ferric oxide (an important
adsorbent).
An example of such conditions is found at a creosote
plume in Pensacola, Florida (Cozzarelli et al.( 1987). Elevated
concentrations of Fe, COZ, and CH4and depleted con-
centrations of dissolved oxygen and NO"3 are associ-
ated with the biodegrading creosote plume. Barium,
molybdenum, manganese, nickel, and strontium are as much
as two orders of magnitude greater than background levels.
187
-------
'K
fMn
"Co
Ru
125Sb 134Cs
144
Ce
152
Eu
Prefilter
450 nm
200 nm
50 nm
3nm
Dissolved
Figure 12-12. Fraction of isotopes in different sized fraction in ground water near the Nevada Test Site (data from Buddemeir and
Hunt, 1988).
Product at
Residual Saturation
\
Biodegradation
Consumption of Oxygen
Consumption of Organic Matter
Production of CO 2
Dissolved Organic
Solutes Entering
Aquifer System
Decreasedpe
DecreasedpH
Groundwater
mmmm
Flow
Natural HFO
High Adsorption Capacity
Adsorbed Metal Ions
Dissolution of HFO
Desorption of Natural Metals
Reduced Sorptive Capacity
Competition for Adsorption Sites
Figure 12-13. Indirect effects of biodegradation of organic contaminants on inorganic constituents.
188
-------
Well A
WellB
(AB)
(AC)
Figure 12-14. Application of mass balance computational
models.
These chemical alterations are found over a much wider area
both horizontally and vertically than the organic contaminants
themselves. Similar conditions also have been described at a
crude oil spill in Bemidji, Minnesota, (Siegel, 1987) where
elevated concentrations of iron, aluminum, and silica are
reported. While many of these elements are not toxic they can
nevertheless pose costly problems of scaling and clogging
during pump-and-treat remediation.
12.4 Computational Tools
The sections above discussed the importance of complex-
ation to understanding the controlling processes of site reme-
diation with inorganic contaminants. Determining the
concentration of each of the complexes is a computationally
complex task that is left to computational algorithms. In
addition, other types of computationally intensive chemical
calculations would be useful, and there are a variety of com-
putational tools to assist in such calculations. In general, these
tools can assist in the calculation of (1) mass balance, (2)
chemical speciation, (3) mass transfer, and (4) multicompo-
nent transport. While some of these algorithms are readily
available at little or no cost, others are still classified as
research tools and are not likely to be available for general use
for several years.
12.4.1 Mass Balance
Mass balance calculations can be applied to a system
such as that illustrated in Figure 12-14. If the chemical com-
position of the water is known at locations A and B along the
flow path, then the change in concentration of each of the
elements along the flow path is known. If the reactions that
take place between the two wells are known, then the amount
of each reaction can be calculated. Reactions such as (1)
mineral dissolution/precipitation, (2) gas exchange, (3) ion
exchange, (4) simple isotope balances, (5) oxidation-reduc-
tion, and (6) the mixing of waters can be included in such
calculations. The code BALANCE (Parkhurst et al, 1982) is a
readily available FORTRAN code that can run on personal
computers. So far this code has been used to study the
geochemical evolution of natural waters (e.g., Plummer and
Back, 1980), yet it has not been applied to the transport of
contaminants or the performance of remediation activities. A
mass balance model such as BALANCE should not be used
by itself but should be used in conjunction with chemical
speciation and mass transfer tools as well as practical knowl-
edge of chemical systems.
12.4.2 Chemical Speciation
Chemical speciation algorithms are used to calculate the
concentration and activities of each of the chemical species
that are in solution. The data requirements for the proper use
of such models include accurate field pH, temperature, and
alkalinity. In addition, a complete inorganic chemical analysis
is required. A complete chemical analysis requires the con-
centration of all of the major anions and cations and the
priority metals and anions under investigation. Some knowl-
edge of the redox conditions within the aquifer is useful,
particularly the total concentration of each of the redox states
of the metals of concern. Most chemical speciation models
also calculate and print out the mineral saturation indices.
There are several chemical speciation models available.
WATEQ4F and SOLMNEQ88 are versions of models that
were originally published in the mid- 1970s by the United
States Geological Survey (e.g., Kharaka et al., 1988; Ballet
al., 1979; Kharaka and Barnes, 1973; Plummer et al., 1976
Truesdell and Jones, 1974). Chemical speciation also is per-
formed by mass transfer models (see below); it maybe more
practical to have a single program for all such calculations.
12.4.3 Mass Transfer
Mass transfer models allow calculation of how much of a
given mineral phase must react for the water to reach equilib-
rium with that phase and achieve the pH and EHof the
equilibrated solution. The basic data requirements for the use
of this type of model are similar to those for chemical specia-
tion models. There are several available mass transfer models,
including PHREEQE (Parkhurst et al., 1980), EQ3/6 (Wolery,
1979, 1983), andMINTEQ (Felmy et al., 1984). MINTEQ has
an extensive data base that includes many of the priority
metals that are of interest at waste sites. MINTEQ is also the
only model that includes choices for adsorption processes,
including (1) ion exchange, (2) Langmuir isotherms, (3)
Freundlich isotherms, (4) double-layer model, (5) Stem-layer
model, and (6) triple-layer model.
12.4.4 Multicomponent Transport
The ultimate tool for assisting in the design of aquifer
remediation strategies is a computational algorithm that ac-
counts for the physical process of advection as well as all of
the chemical processes discussed above. While progress has
been made in this area (Jennings et al., 1982; Yeh and
Tripathi, 1989), these models are not generally available.
Although these models are still considered to be research
tools, there is much work being done to complete models that
soon will be available for general use.
189
-------
12.5 References
Ball, J.W., E.A. Jenne, and D.K Nordstrom. 1979. WATEQ2
— A Computerized Chemical Model for Trace and Major
Element Speciation and Mineral Equilibria of Natural
Waters. In: Chemical Modeling in Aqueous Systems:
Speciation, Sorption, Volubility, and Kinetics, E.A. Jenne
(ed.), ACS Symp. Series 93, American Chemical Society,
Washington, DC, pp. 815-836.
Ball, J. W., D.K. Nordstrom, and E.A. Jenne. 1980. Additional
and Revised Thermochemical Data and Computer Code
for WATEQ2: A Computerized Chemical Model for
Trace and Major Element Speciation and Mineral Equi-
libria of Natural Waters. U.S. Geological Survey Water
Resources Investigations No. 78-116.
Benjamin, M.M. and J.O. Leckie. 1981. Multiple-Site Ad-
sorption of Cd, CU. Zn, and Pb on Amorphous Iron
Oxyhydroxides. J. Coll. Interface Sci. 79(2):209-221.
Black, J.A. and J.H. Heil. 1982. Municipal Solid Waste
Leachate and Scavengerwaste: Problems and Prospects in
Brookhaven Town. In: Proceedings of the Northeast Con-
ference on the Impact of Waste Storage and Disposal on
Ground-Water Resources, R.P. Voitski and G. Levine
(eds.), U.S. Geological Survey and Cornell University,
2.1:1-12.
Booz-Allen and Hamilton, Inc. 1987. ROD (Record of Deci-
sion) Annual Report FY 1986. U.S. Environmental Pro-
tection Agency (NTIS PB87-199550), 182 pp.
Brookins, D.G. 1988. Eh-pH Diagrams for Geochemistry.
Springer-Verlag, New York, 176 pp.
Buddemeir, R.W. and J.R. Hunt. 1988. Transport of Colloidal
Contaminants in Groundwater Radionuclide Migration
at the Nevada Test Site. Applied Geochemistry 3:535-
548.
Cook, D.K. and R.G. DiNitto. 1982. Evaluation of Groundwa-
ter Quality in East and North Wobum, Massachusetts. In:
Proceedings of the Northeast Conference on the Impact
of Waste Storage and Disposal on Ground-Water Re-
sources, R.P. Voitski and G. Levine (eds.), U.S. Geologi-
cal Survey and Cornell University, 4.2:1-20.
Cozzarelli, I.M., M.J. Baedecker, and J.A. Hopple. 1987.
Effects of Creosote Products on the Aqueous Geochemis-
try of Unstable Constituents in a Surficial Aquifer. In:
U.S. Geological Survey Program on Toxic Waste-
Ground-Water Contamination: Proceeding of the Third
Technical Meeting, Pensacola, Florida, March 23-27,
1987, B.J. Franks (ed.), U.S. Geological Survey Open-
File Report 87-109, pp. A15-A16.
Drever, J.I. 1989. The Geochemistry of Natural Waters, 2nd
ed. Prentice-Hall, Englewood Cliffs, NJ. [First edition
1982].
Dzombak, D.M. 1986. Towards a Uniform Model for Sorp-
tion of Inorganic Ions Hydrous Oxides. PhD Dissertation,
Department of Civil Engineering, Massachusetts Institute
of Technology.
Dzombak, D.A. and F.M.M. Morel. 1986. Sorption of Cad-
mium on Hydrous Ferric Oxide at High Sorbate/Sorbent
Ratios: Equilibrium, Kinetics, and Modelling. J. Colloid
Interface Sci. 112 2):588-598.
Dzombak, D.A. and F.M.M. Morel. 1990. Surface Complex-
ation Modelling, Hydrous Ferric Oxide. John Wiley &
Sons, New York, 393 pp.
Dzombak, D.A., W. Fish, and F.M.M. Morel. 1986. Metal-
Humate Interaction. 1. Discrete Ligand and Continuous
Distribution Models. Environ. Sci. Technol. 20:669-675.
Evans, D. W., J.J. Alberts, and R.A. Clark. 1983. Reversible
Ion-Exchange of Cesium-137 Leading to Mobilization
from Reservoir Sediments. Geochimica et Cosmochimica
Acta 47(11): 1041 -1049.
Felmy, A.R., D.C. Girvin, E.A. Jenne. 1984. MINTEQ A
Computer Program for Calculating Aqueous Gemchemi-
cal Equilibria. EPA/600/3-84-032 (NTIS PB84-157148).
Fish, W. 1990. Subsurface Redox Chemistry: A comparison
of Equilibrium and Reaction-Based Approaches. In: Metal
Speciation in Groundwater, H. Allen E.M. Perdue, and D.
Brown (eds.), Lewis Publishers, Chelsea, MI.
Fish, W. (in press). Subsurface Transport of Gasoline-Derived
Lead. Ground Water.
Fish, W. and M.S. Elovitz. 1990. Redox and Solvation Inter-
actions between Hexavalent Chromium and Hydroxy-
lated Organic Compounds. U.S. EPA Contract Report
90-R-8 14136-01-0.
Garrels, R.M. and C.L. Christ. 1965. Solutions, Minerals, and
Equilibria. Harper and Row, New York, 450 pp.
Gschwend, P.M. and M.D. Reynolds. 1987. Monodisperse
Ferrous Phosphate Colloids in an Anoxic Groundwater
Plume. J. Contaminant Hydrology 1:309-327.
Hanzlik, R.P. 1976. Inorganic Aspects of Biological and
Organic Chemistry. Academic Press, New York.
Helfferich, F. 1962. Ion Exchange. McGraw-Hill, New York.
Honeyman, B. D., K.F. Hayes, and J.O. Leckie. 1982. Aque-
ous Chemistry of As, B, Cr, Se, and V with Particulm
Reference to Fly-ash Transport Water. EPRI-910-1. Elec-
tric Power Research Institute, Palo Alto, California.
Jennings, A. A., D.J. Kirkner, and T.L. Theis. 1982. Multi-
component Equilibrium Chemistry in Ground Water Qual-
ity Models. Water Resources Reach 18: 1089 -1096.
190
-------
Kharaka, Y.K. and I. Barnes. 1973. SOLMINEQ: Solution-
Mineral Equilibrium Computations. U.S. Geological Sur-
vey, Menlo Park, CA (NTIS PB215-899).
Kharaka, Y.K., W.D. Gunter, P.K. Aggarwal, E.H. Perkins,
and J.D. DeBraal. 1988. SOLMINEQ.88: A Computer
Program for Geochemical Modeling of Water-Rock In-
teractions. U.S. Geological Survey Water-Resources In-
vestigations Report 88-4227, 420 pp.
Laha, S. and R.G. Luthy. 1990. Oxidation of Aniline and
Other Primary Aromatic Amines by Manganese Dioxide.
Environ. Sci. Technol. 24:363-373.
Lmdberg, R.D., and D.D. Runnels. 1984. Ground Water Re-
dox Reactions: An Analysis of Equilibrium State Applied
to Eh Measurements and Geochemical Modeling. Sci-
ence 225:925-927.
McDowell-Boyer, J.R. Hunt, and N. Sitar. 1986. Particle
Transport Through Porous Media. Water Resources Re-
search 22:1901-1921.
Moore, J.W. and S. Ramamoorthy. 1984. Heavy Metals in
Natural Waters. Springer-Verlag, New York, 268 pp.
Morel, F.M.M. 1983. Principles of Aquatic Chemistry. Wiley
Interscience, New York.
Murarka, IP. and D.A. Mclntosh. 1987. Solid-Waste Envi-
ronmental Studies (SWES): Description, Status, and Avail-
able Results. EPRIEA-5322-SR. Electric Power Research
Institute, Palo Alto, CA.
Nordstrom, O.K. and J.L. Munoz. 1985. Geochemical Ther-
modynamics. Benjamin/Cummings Publishing, Menlo
Park, CA, 477 pp.
Palmer, C. D., W. Fish, and J.F. Keely. 1988. Inorganic Con-
taminants: Recognizing the Problem. In: Proc. Second
Nat. Outdcmr Action Conf. on Aquifer Restoration, Ground
Water Monitoring, and Geophysical Methods, National
Water Well Association, Dublin, OH, pp. 555-579.
Parkhurst, D.L., D.C. Thorstenson, and L.N. Plummer. 1980.
PHREEQ — A Computer Program for Geochemical Cal-
culations. U.S. Geological Survey Water-Resources In-
vestigations Report 76-13, 81 pp.
Parkhurst, D.L., L.N. Plummer, and D.C. Thorstenson. 1982.
BALANCE-A Computer Program for Calculating Mass
Transfer for Geochemical Reactions in Ground Water.
U.S. Geological Survey Water-Resources Investigations
Report 82-14 (NTIS PB82-255902), 29 pp.
Plummer, D.L. and W. Back. 1980. The Mass Balance Ap-
proach: Application to Interpreting the Chemical Evolu-
tion of Hydrologic Systems. Am. J. Science 280:130-142.
Plummer, L.N., B.F. Jones, and A.H. Truesdell. 1976.
WATEQF — A FORTRAN IV Version of WATEQ, a
Computer program for Calculating Chemical Equilib-
rium of Natural Waters. U.S. Geological Survey Water-
Resources Investigations Report 75-61, 73pp.
Rees, T.F. 1987. A Review of Light-Scattering Techniques
for the Study of Colloids in Natural Waters. J. Contami-
nant Hydrology 1:431-439.
Reichenbcrg, D. 1966. Ion Exchange Selectivity. In: Ion
Exchange and Solvent Extraction, Vol. 1, J.A Marinsky
(ed.), Marcel Dekker, New York, pp. 227-276.
Schindler, P.W. 1981. Surface Complexes at Oxide-Water
Interfaces. In: Adsorption of Inorganic at Solid-Liquid
Interfaces, M.A. Anderson and A.J. Rubin (eds.), Ann
Arbor Science, Ann Arbor, MI, pp. 1 -49.
Schindler, P.W. and W. Stumm. 1987. The Surface Chemistry
of Oxides, Hydroxides, and Oxide Minerals. In: Aquatic
Surface Chemistry, W. Stumm (ed.), John Wiley & Sons,
New York, pp. 83-110.
Sholkovitz, E.R. 1985. Redox-related Geochemistry in Lakes:
Alkali Metals, Alkaline Earth Metals, and Cesium-137,
In: Chemical Processes in Lakes, W. Stumm (ed.), Wiley-
Interscience, New York.
Siegel, D.I. 1987. Geochemical Facies and Mineral Dissolu-
tion, Bemidji, Minnesota, Research Site. In: U.S. Geo-
logical Survey Program on Toxic Waste-Ground-Water
Contamination: Proceeding of the Third Technical Meet-
ing, Pensacola, Florida, March 23-27, 1987, B.J. Franks
(ed.), U.S. Geological Survey Open-File Report 87-109,
pp. C13-C15.
Smith, R.L. and A.E. Marten. 1976. Critical Stability Con-
stants. Plenum, New York.
Sposito, G. 1984. The Surface Chemistry of Soils. Oxford
University Press, New York.
Sposito, G. 1986. Sorption of Trace Metals by Humic Materi-
als in Soils and Natural Waters. CRC Critical Reviews in
Environmental Control 16:193-229.
Stone, A.T. 1986. Adsorption of Organic Reductants and
Subsequent Electron Transfer on Metal Oxide Surfaces.
In: Geochemical Processes at Mineral Surfaces, J.A. Davis
and K.F. Hayes (eds.s), ACS Symp. Series 323, Ameri-
can Chemical Society, Washington, DC, pp. 446-461.
Stumm, W. and J.J. Morgan. 1981. Aquatic Chemistry, 2nd
ed. Wiley Interscience, New York.
Stumm, W., H. Hohl, F. Dalang. 1976. Interaction of Metal
Ions with Hydrous Oxide Surfaces. Croat. Chem. Acts.
48(4):491-504.
Truesdell, A.H. and B.F. Jones. 1974. WATEQ, A Computer
Program for Calculating Chemical Equilibria of Natural
Waters. J. Research U.S. Geological Survey 2:233-248.
191
-------
Vanselow, A.P. 1932. Equilibria of the Base Exchange Reac- Wolery, T.J. 1979. Calculation of Chemical Equilibrium Be-
tions of Bentonites, Permutites, Soil Colloids and Zeo- tween Aqueous Solutions and Minerals: The EQ3/6 Soft-
lites. Soil Science 33:95-113. ware Package. Report UCRL 52658. Lawrence Livermore
National Laboratory, Livermore, CA.
Voudrias, E.A. and M. Reinhard. 1986. Abiotic Organic Re-
actions at Mineral Surfaces. In: Geochemical Processes Wolery, P.J. 1983. EQ3NR, a Computer Program for Geo-
at Mineral Surfaces, J.A. Davis and K.F. Hayes (eds.), chemical Aqueous Speciation-Volubility Calculations.
ACS Symp. Series 323, American Chemical Society, Report UCRL 53414. Lawrence Livermore National Labo-
Washington, DC, pp. 462-486. ratory, Livermore CA.
Westall, J.C. and H. Hohl. 1980. A Comparison of Electro- Yeh, G.T. and V.S. Tripathi. 1989. A Critical Evaluation of
static Models for the Oxide/Solution Interface. Advances Recent Developments in Hydrogeochemical Transport
Coll. Interface Sci. 12(2):265-294. Models of Reactive Multichemical Components. Water
Resources Research 25:93-108.
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Chapter 13
Characterization of Subsurface Degradation Processes
J. Michael Henson
When chemical constituents enter the subsurface envi-
ronment, they are subjected to physical, chemical, and bio-
logical processes that ultimately determine their fate and
transport characteristics. Knowledge of the degradation pro-
cesses that determine the fate of organic compounds in the
subsurface can be used to guide remediation efforts at sites
that have been affected. The physical processes that control
the transport of constituents in the subsurface are discussed in
previous chapters, This chapter describes biological and
nonbiological processes that may control the fate of organic
chemicals once they have entered the subsurface. An under-
standing of these principles will aid in the efficient and cost-
effective remediation of releases of organic constituents.
Objectives of this chapter are to:
• Present information about abiotic degradation pro-
cesses.
• Present information about biological degradation pro-
cesses.
• Provide a basis for site evaluation to determine the
potential for biological remediation.
• Build the foundation for the discussion of bioreme-
diation of soils (Section 15.2.2) and ground water
(Section 16.3).
This chapter discusses two classes of transformations that
may occur in the subsurface-abiotic and biologic transfor-
mation. Abiotic reactions are those reactions that do not
involve metabolically active organisms, a product of a living
cell, or a product of a previously living organism. Some
examples of products of cells are extracellular enzymes, he-
moprotein, iron porphyrins, cytochromes, flavins, and re-
duced pyridine nucleotides.
13.1 Abiotic Transformation Reactions
Hydrolysis, substitution, elimination, and oxidation-re-
duction are the abiotic reactions that will be discussed in this
chapter. These reactions produce a variety of end-products
whose presence may play a role in decisions made to select
compounds for the remedial investigation phase. The results
of an abiotic reaction may enhance the biological degradability
of a compound and provide possible treatment of the parent
compound. Dragun (1988) provides an excellent presentation
of abiotic reactions.
13.1.1 Hydrolysis
Hydrolysis reactions are those reactions where an organic
chemical reacts with either water or a hydroxide ion to pro-
duce an alcohol. The following equations represent these
reactions:
R-X + H.O—> R-OH + H++ X
R-X + OH—> R-OH + X
In these reactions, either H20 or OH act as a nucleophile and
attack the electrophile, RX, to displace the leaving group, X.
This type of reaction is referred to as a nucleophilic displace-
ment reaction and in this example results in the formation of a
daughter product that is an alcohol. For a more detailed
discussion of this nucleophilic displacement reaction mecha-
nism, see Dragun (1988). The rate of hydrolysis reactions is
typically first order with respect to the concentration of the
compound. The rate of a first-order reaction increases as the
concentration of the organic compound increases. The first-
order rate constant k can be calculated as:
k = (2.303/t) log[C0/(C0- C,)]
where t is time, C0is initial concentration, and C,is concentra-
tion at t.
The time required for half of the concentration of the com-
pound to degrade is known as the half-life, t,/2, and is calcu-
lated as:
\= 0.693/k
Some examples of hydrolysis half-lives for some organic
compounds are presented in Table 13-1. A more extensive
listing of hydrolysis half-lives can be found in Dragun (1988).
The rates of hydrolysis vary from compound to com-
pound and can be on the order of hours to years. The rates of
hydrolysis also indicate the susceptibility of the compounds to
hydrolysis. Some examples of organic chemicals that are
subject to hydrolysis are alkyl halides, carbamates, chlori-
nated amides, esters, and epoxides. Examples of chemicals
193
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Table 13-1. Selected Hydrolysis Half-Lives for a Variety of
Organic Compounds
Organic Compound
Atrazine
Chloroethane
Chloromethane
Diazinon
Dichloromethane
Epoxyethane
Ethyl acetate
lodoethane
Malathion
Methyl parathion
Parathion
Tetrachloromethane
Trichloromethane
Trimethylphosphate
Hydrolysis Half-Life (pH = 7)
2.5 hours
38 days
339 days
9.5 days (pH = 6)
704 years
12 days
136 days (pH = 6)
49 days
8.1 days (pH = 6)
10.9 days (pH = 6;
17days(pH = 6)
7000 years
3500 years
1.2years
that are more resistant to hydrolysis are aldehydes, alkanes,
alkenes, and compounds with carboxy - or nitro-substituents.
Once an organic compound enters the subsurface, envi-
ronmental factors can decrease or increase the hydrolysis half-
life that might be projected from the results of a laboratory
evaluation. One effect that soil can have on the hydrolysis
half-life is on localized pH differences. The pH at the surface
of the soil particles can be very different from the overall soil
pH. These localized effects may alter the half-life by enhanc-
ing or inhibiting the hydrolysis reaction. Another effect of soil
on half-life results from metal ions that are present as normal
components of the soil. These metals can serve as catalysts for
organic reactions. A third environmental factor is adsorption
of the organic compound to the soil particles, which can affect
the rate of hydrolysis reactions. By adsorbing to the soil
particle, the compound is in effect removed from the water.
Other factors such as soil water content and the type of soil
matrix also can affect the rate of hydrolysis.
13.1.2 Substitution
Hydrolysis reactions are classified as a type of substitu-
tion reaction but they are presented first because of the
predominance of water, which causes the reactions to occur.
Other chemicals in the subsurface can cause substitution
reactions to occur. An example of a substitution reaction
involves hydrogen sulfide acting as the nucleophilic agent to
attack organic compounds, which result in the production of
sulfur-containing compounds.
13.1.3 Elimination
Elimination reactions cause the loss of two adjacent
groups from within the molecule resulting in the formation of
a double bond. The reaction occurs as:
R-CHX.-CH^ -
R-CH=CH2 + X, + X2
One example of an elimination reaction is the formation
of 1,1 -dichloroethene (1,1-DCE) from 1,1,1-trichloroethane
(1,1,1-TCA). An additional formation product of an abiotic
reaction was the detection of acetic acid formed as a result of
substitution. The ratio of acetic acid to 1, 1-DCE was about 3:1
(Cline et al, 1988). Elimination also can result in the forma-
tion of bromoethene from 1,2-dibromoethane and
bromopropene from 1,2-dibromopropane (Dragun, 1988).
13.1.4 Oxidation-Reduction
Oxidation is the net removal of electrons from an organic
compound, while reduction is the net gain of electrons by an
organic compound. These reactions are coupled by the trans-
fer of electrons from one compound to another. The oxida-
tion-reduction couples in soil systems are complex and multiple.
In many instances, if a biological response to an organic
compound occurs, the biological system will tend to become
predominant. Inorganic redox reactions are discussed further
in Section 12.1.3.
Abiotic reactions may occur in the subsurface by a vari-
ety of mechanisms and at varying rates. The use of abiotic
reactions as a remediation technology has not received a lot of
attention, but may provide an alternative treatment in some
instances. Abiotic reactions may occur in conjunction with
biological reactions and make some compounds more suscep-
tible to biodegradation. Abiotic reactions may not always
provide extensive treatment of the organic compound but the
treatment that does occur may produce a compound of less
environmental concern.
13.2 Microbiological Transformations in the
Subsurface
Microbiological transformations are the second class of
processes that have an impact on the fate of organic com-
pounds once they enter the subsurface. This class of processes
can result in either partial or complete degradation of the
organic compounds to detoxify or remove them from the
subsurface. The knowledge of biological responses to various
organic compounds can be utilized during the site investiga-
tion process to collect data that will aid in evaluating potential
remediation alternatives. These remediation alternatives can
include biological remediation.
When addressing biological transformations, biodegrada-
tion is typically used to mean complete degradation. How-
ever, biodegradation specifically refers to the biological
transformation of an organic compound without regard to the
extent of transformation. Mineralization specifically refers to
the conversion of an organic compound to carbon dioxide (or
methane in anaerobic environments), water, and a halogen
atom, if the parent compound was halogenated.
Knowledge of biological responses to organic compounds
that may occur under different microbial growth conditions
provides an understanding of the metabolic potential by which
microorganisms may transform these compounds. For ex-
ample, if partial degradation of an organic compound occurs,
the daughter products formed may or may not be of environ-
mental concern. The observation of microbial intermediates
194
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of metabolism indicates that a biological response to the
parent compounds has occurred and that the potential for site
remediation through biological processes exists.
The use of microorganisms for remediation of sites af-
fected with organic compounds is gaining increasing interest.
This process of bioremediation requires an integrated ap-
proach involving the disciplines of microbiology, hydrogeol-
ogy, and engineering. The relationship of these three disciplines
is analogous to a "three-legged stool," in that if one of the legs
is weak, the stool cannot support much weight. These three
disciplines also must be augmented with an awareness of the
principles of the physical-chemical interactions that are the
subject of previous chapters, and an understanding of the
regulatory requirements in which the application of bioreme-
diation will take place. This section will provide a basic
understanding of the principles of microbial ecology as re-
lated to the subsurface. This understanding can aid in the
evaluating of sites affected with organic compounds and
provide a basis for the following chapters where examples of
bioremediation will be presented.
13.2.1 Microbial Ecology of the Subsurface
Although it is now known that significant numbers of
microorganisms are distributed throughout the subsurface
(Back, 1989; Ghiorse and Wilson, 1988), it was once sug-
gested that numbers of microorganisms in soil decreased with
depth (Waksman, 1916). More recent investigations, how-
ever, routinely detect microorganisms in aquifers. These in-
vestigations include aquifers not known to have been affected
with organic compounds and aquifers that have received
inputs of organic compounds.
The development of techniques to investigate water table
aquifers was instrumental to the elucidation of the microbial
ecology of the subsurface. McNabb and Mallard (1984) de-
scribed sampling techniques designed to prevent the micro-
bial contamination of samples retrieved from the terrestrial
subsurface. These techniques rely on the collection of cores
from the depth to be investigated. After collection, the outer-
most layer can be removed in the field with alcohol-sterilized
devices designed to strip away the soils that were in contact
with drilling equipment. These techniques produce a subcore
of the original core in an aseptic manner. Subcores can be
obtained in the laboratory by a variety of mechanisms, as long
as aseptic techniques are used. For field or laboratory condi-
tions, the subcore can be collected under anaerobic (Beeman
and Suflita, 1987) as well as aerobic conditions.
The collection of subsurface samples using aseptic tech-
niques to prevent intrusion of microorganisms not representa-
tive of the subsurface has yielded considerable information
about the microbial ecology of the subsurface. For example,
Wilson et al. (1983) and Balkwill and Ghiorse (1985) reported
the presence of between 1 and 10 million microorganisms per
gram of sediment using the Acridine Orange Direct Count
(AODC) staining technique to count the microorganisms. The
same authors, using a plate count assay to count viable micro-
organisms, detected between 200,000 and 2.5 million micro-
organisms per gram of sediment in two aquifers that were not
known to have received input of organic compounds. Similar
ranges of counts for microorganisms for shallow aquifers not
receiving organic chemicals arc shown in Table 13-2.
Beeman and Suflita (1987) reported a range of 11 to 17x
106 cells (g dry wgt) measured by AODC in a sand aquifer
receiving landfill leachate in Norman, Oklahoma. Similar
ranges of microbial counts by AODC were observed by Erlich
et al. (1983) and Webster et al. (1985) for two different
aquifers that were affected with creosote compounds.
The results of these investigations indicate that the terres-
trial subsurface whether pristine or not is populated by micro-
organisms. These numbers of microorganisms are relatively
high and were detected in a variety of geologic environments
and depths. Analysis of subsurface samples indicates that the
microorganisms are predominantly attached to the subsurface
soil particles (Harvey et al., 1984). Evidence also is accumu-
lating that even deeper geologic environments are inhabited
by microorganisms (Updegraff, 1982).
Biochemical diversity of microorganisms present in the
subsurface is evidenced by the variety of organic compounds
reported to be metabolized. Petroleum hydrocarbons, includ-
ing fuels, creosote constituents, and products of coal gasifica-
tion, are reported to be substrates for subsurface
microorganisms under a variety of growth conditions. Table
13-3 presents examples of organic compounds metabolized
by subsurface microorganisms.
Table 13-2. Microbial Cell Counts for Selected Aquifers That Were Not Receiving Known Inputs of Organic Compounds
Study Site
Lula, OK
Pickett, OK
Aquifer Type Sample Depth (m)
Sand and Gra vel 5
Sand 5.5
Total Count x 10s
Cells g Dry Wgt
3.8 to 9.3
5.2
References
Balkwill and Ghiorse, 1985;
Wilson, et al. 1983
Balkwill and Ghiorse, 1985
Ghiorse and Balkwill, 1985
Fort Polk, LA
Dayton, OH
Alberta, Can.
Loamy Clay
Gravel
Marmot Basin
5
10-12
1.5
9.8
0.036 to 0.06
0.05 to 2.5
Ghiorse and Balkwill, 1983
Ventullo and Larson, 1985
Laddetal., 1982
195
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13.2.2 Relationship of Environmental Factors to
Biodegradation
Microorganisms require a suitable set of environmental
factors in order to grow. These factors include the chemical
and physical parameters of pH, available water or osmotic
pressure, temperature, and absence of toxic conditions.
The pH of the environment is an easily measured param-
eter and indicates the potential for microbial activity. Many
microoganisms grow best in the pH range of 6 to 8. Microbial
life at extremes of pH does occur and, therefore, a pH outside
of the 6 to 8 range does not exclude microbial growth. Growth
of microorganisms can raise or lower the pH by producing
end-products that affect pH or by removing the parent com-
pounds, thus, affecting the pH. The measurement of pH in
ground water could indicate the potential for microorganisms
to grow in the aquifer.
Temperature generally affects microbial growth in that an
increase in temperature results in an increase in microbiologi-
cal growth. Microorganisms have lower, upper, and optimum
temperature limits for growth. Many microorganisms in the
soil have an optimum temperature for growth between 10°
and 30°C. Temperatures of ground waters within the United
States are within this range (Dragun, 1988).
Microorganisms require water for active growth. The
availability of water depends on the number of molecules
present in the solution. An increase in the number of mol-
ecules, relative to the number of molecules within the micro-
bial cells, results in the movement of water from the cell into
the surrounding environment as a result of osmosis. The
opposite effect occurs when the number of molecules outside
the microbial cell is less than the number inside the microbial
cell. The soil moisture content is sometimes critical to the
growth of microorganisms. If the moisture content is near
saturation, transfer of oxygen may become a growth-limiting
factor. If the soil is dry, growth of the microorganisms will be
very limited.
13.2.3 Microbial Metabolism
The ability of communities of microorganisms to me-
tabolize many types of organic compounds including syn-
thetic organic compounds is well documented (Alexander,
1981; Gibson, 1984). A number of these organic compounds
are utilized by microorganisms as a source of carbon and
energy. The degradation of the compounds may not occur at
the initial time of release to the environment. A period of time
may elapse before an increase in the rate of degradation is
observed. This period of time is referred to as an adaptation or
acclimation period. The adaptation period may vary with the
compound and the environmental conditions into which the
compound is released. For example, under anaerobic condi-
tions, adaptation periods may be as long as several months.
Once the adaptation occurs, however, the rate of degradation
becomes a function of the processes controlling the availabil-
ity of nutrients to the microorganisms and not of an intrinsic
metabolic property of the microorganisms. In addition, once
the microbial community adapts to a particular organic com-
pound or compounds, the compound or compounds can con-
tinue to be added without re-adaptation. The microbial
community, thus, becomes enriched in members that can
metabolize the organic compounds.
An additional opportunity for microbial degradation is
through a process of nongrowth metabolism. In this process,
the microorganisms do not use the organic compound as a
source of carbon and energy, which results in growth. Instead,
the microorganisms cometabolize a substance that cannot be
utilized for growth in the presence of a compound that can be
utilized for growth. The cometabolized compound is often
transformed into an intermediate that can undergo transforma-
tion by other microorganisms. A specific example, to be
discussed in more detail later, is the degradation of trichloro-
ethene and dichloroethene by microorganisms that are grow-
ing on methane and fortuitously react with the halogenated
compounds.
The ability of microorganisms to degrade organic com-
pounds depends on the presence of a terminal electron accep-
tor (TEA), as well as other nutrients. The TEA receives
electrons from a series of oxidation-reduction reactions within
the cell that generate energy allowing the microorganism to
grow. Some microorganisms can use several TEAs whereas
other microorganisms can use only one. If more than one TEA
is present when an organic compound enters the environment,
the one that results in the highest energy transfer will be used
first. Next, the TEA with the second highest energy transfer
will be used, and so on until either the organic compound is
removed or the TEAs have been consumed.
Table 13-3. Representative Examples of the Diversity of Organic Compounds Metabolized by Subsurface Microorganisms
Organic Compound Metabolized Growth Conditions
References
Petroleum Hydrocarbons
Hydrocarbons
Creosote/Coal Gas Compounds
Creosote Compounds
Aerobic
Anaerobic
Aerobic
Anaerobic
Ehrlich,etal., 1985; Jamison, et al., 1975; Lee, etal., 1988; Lee and Ward,
1985; Raymond, etal., 1976; Wilson and Ward, 1987; Wilson, etal., 1985b
Grbic-Galic and Vogel, 1987; Vogel and Grbic-Galic, 1986;
Wilson and Rees, 1985
Humenick, etal., 1982; McGinnis, etal., 1988; Wilson, etal., 1985a
Erlich, et al., 1983; Smolensk! and Suflita, 1987
Dayton, OH
Alberta, Can.
Gravel
Marmot Basin
10-12
1.5
0.036 to 0.06
0.05 to 2.5
Ventullo and Larson, 1985
Laddetal., 1982
195
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Table 13-4 presents the relative energy charge associated
with the consumption of various TEAs. The succession of
metabolic events will proceed from the reactions with TEAs
that can transfer the most energy, which are denoted by the
most negative values in Table 13-4. The succession also is
related to the toxicity of TEAs for groups of bacteria. For
example, methanogenic bacteria are inhibited by oxygen;
therefore, the development of active methanogenesis usually
does not occur until oxygen is removed from the environment
and reducing conditions are established.
Aerobic respiration is the process of consuming organic
compounds with oxygen serving as the TEA. The end-product
of the respiration of oxygen is water. The degradation of
hydrocarbons also requires oxygen as a cosubstrate where the
oxygen is inserted into the hydrocarbon molecule.
Once the metabolic demand for oxygen exceeds the rate
of supply, the anoxic conditions that are established allow
other TEAs to be utilized. The next TEA that, if present,
would be used is nitrate. The respiration of nitrate is referred
to as denitrification and results in the production of nitrogen
gas (Knowles, 1982). The transfer of energy is similar to that
of the respiration of oxygen. Many of the organisms that use
nitrate as an electron acceptor also use oxygen so that accli-
mation of a new population of microorganisms may not be
required.
Once the nitrate has been consumed and the oxidation-
reduction state becomes reducing, the respiration of sulfate
can begin as a process known as sulfate reduction (Postgate,
1979). Sulfate reduction results in the production of hydrogen
sulfide, which can be corrosive to equipment and potentially
toxic to humans. Sulfate reduction does not yield as much
energy, only about one-fourth, as does the respiration of
nitrate or oxygen.
If nitrate is present in a reducing environment, its respira-
tion does not result in the production of nitrogen gas; instead,
ammonia is produced (Caskey and Tiedje, 1980). The respira-
tion of nitrate under reducing conditions does not transfer as
much energy as the respiration of oxygen.
As the conditions become more reducing and alternative
TEAs are consumed, the respiration of carbonate will result in
the production of methane. The microorganisms that carry out
Table 13-4. Comparison of Free Energy Values for Metabolism
of Glucose in the Presence of Various Electron
Acceptors
Equation
Glucose + Nitrite -
Glucose + Oxygen
Glucose + Nitrate -
Glucose + Sulfate- -
Glucose + COf - ->
Glucose + Glucose-
- -> CO, + Hp + N2
---> co, + Hp
- -> COt + H2O + Nitrite
-> CO2 + Hp + Sulfide
CO., + H2O + Methane
- -> CO2 + Ethanol
kcal/Electron
Equivalent
-32.3
-28.7
-19.4
-4.9
-4.3
-2.4
this reaction are known as methanogenic bacteria. The energy
transferred during methane production is about one-fourth
that of the respiration of oxygen or nitrate.
The ability of microorganisms to carry out a variety of
respirations provides the opportunity to collect data during the
site investigation phase that indicate whether a microbiologi-
cal response to organic compounds has occurred. If accurate
measurements of dissolved oxygen (DO) in ground water
indicate that oxygen is present outside a plume of organic
compounds and DO is not detected within the plume, then a
biological response may have occurred to consume the oxy-
gen. If methane or another of the respiratory end-products is
detected within the plume, the results suggest that a biological
response has occurred and that reducing conditions may exist.
In addition, the range of metabolic capabilities of micro-
organisms extends beyond the respiration of oxygen. Nitrate
is more water soluble than oxygen and may be less costly to
use in the treatment of some affected aquifers. Reducing
conditions allow some biological transformations to occur
that do not occur under oxidizing conditions. One example of
this type of biological transformation is reductive dechlorina-
tion. This microbiological process removes chlorines from
chlorinated compounds (discussed in Section 13.3.2).
13.2.4 Biological Reaction Kinetics
The rate at which microorganisms can remove organic
compounds from the subsurface can be expressed mathemati-
cally to approximate the time required for remediation. The
first-order rate constant is based on the observation that as the
concentration of the organic compound increases, the rate of
degradation increases. The first-order rate constant k is calcu-
lated as:
k = (2.303/t) log[C0/(C0- C,)]
where t is time, C0is initial concentration, and C,is concentra-
tion at t. The time required for half of the concentration of the
compound to degrade is known as the half-life, t/2, and is
calculated as follows:
t,/2= 0.693/k
However, metabolism in microorganisms occurs via enzymes
that become saturated; the substrates are degraded when the
concentration of the substrate continues to increase. Once the
enzymes become saturated, the rate of degradation cannot
increase and the degradation rate curve becomes hyperbolic.
The use of the first-order rate kinetics provides a general
expression of the rate of biodegradation for many compounds.
Dragun (1988) provides a compilation of first-order degrada-
tion rates that should not be used without comparing the
environments from which these samples were taken. A direct
extrapolation of results obtained from one environment to
another environment is typically not useful.
Based on data from McCarty, 1975
197
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13.3 Bioremediation of Organic Compounds in
the Subsurface
13.3.1 General Considerations
The basic premises of microbial ecology are related to
bioremediation in that many organic compounds can be used
by microorganisms as a source of carbon and energy. Many of
the compounds that are considered hazardous can be degraded
in the subsurface if the concentrations are not toxic to the
microorganisms and the appropriate environmental param-
eters can be established. Bioremediation is based on the
understanding of the carbon cycle and extrapolation of com-
pound mineralization in other environments to the subsurface.
Environmental factors, such as pH, oxidation-reduction po-
tential, and temperature, may play a role in determining the
potential for bioremediation. However, the rate at which
nutrients, especially a TEA, can be delivered to the microor-
ganisms may determine whether bioremediation is feasible.
There are several reviews that provide detailed discussions of
bioremediation (Lee et al., 1988; Thomas and Ward, 1989;
Wilson et al., 1986; and Wilson and Ward, 1987).
Certain information is required before design of the bio-
remediation system can begin. An assessment of the site to
evaluate history, geology, and hydrology can provide infor-
mation valuable for bioremediation design. The delivery of
nutrients to subsurface microorganisms for in situ remediation
is dependent on the site hydrology. Sites with low permeabil-
ity, such as those with clays, may not allow the delivery of
nutrients in an efficient manner.
A thorough laboratory assessment of the microbiology
also provides information to indicate whether bioremediation
is an appropriate treatment technology. Some components of
this assessment are:
Evaluate the presence of requisite microorganisms.
Assess potential toxicity to the microorganisms.
• Evaluate nutrient requirements to enhance degrada-
tion activity.
Evaluate the compatibility of the site geochemistry
with the nutrient solution proposed for addition.
Requisite microorganisms are the ones that are capable of
degrading the organic compounds present at the site. For
many sites, these microorganisms are naturally occurring and
just need some nutrients to stimulate their growth. The pres-
ence of these microorganisms at the site is evaluated in
samples representative of the environment to be remediated.
If the remediation is an in situ aquifer remediation, then the
samples should be collected from the aquifer. The microor-
ganisms are predominantly attached to the soils; therefore,
samples of the soils below the water table should be collected.
Several methods exist for collecting the samples. Principles
for collection are discussed in McNabb and Mallard (1984).
Microorganisms present in the samples should be enu-
merated in a manner to indicate the presence of viable micro-
organisms. Staining techniques exist, such as the AODC, but
this technique is limited because it does not indicate the
viability of the microorganisms. The results of viable counts
suggest the environment that was sampled was not so toxic as
to completely inhibit the presence of microorganisms. Tech-
niques such as standard plate counts can be used to detect the
number of general microorganisms present. Plate counts using
a microbial medium containing the compound of interest also
can be used to enumerate the bacteria present in the sample
capable of growth on that compound. The numbers can be
compared before and after treatment to assess whether the
treatment resulted in an increase in the number of microbes
capable of growth on the compound of interest. An increase in
the observed number of bacteria would suggest an effective
process.
The nutrients required to enhance microbial growth are
assessed primarily on the nitrogen and phosphorous require-
ments of the microorganisms. However, the microorganisms
may require other nutrients such as potassium, magnesium,
manganese, and iron. The site's geochemistry may provide
many of these necessary nutrients. The nutrient solution se-
lected should be compatible with the geochemistry of the site
to prevent possible precipitation of minerals, which might
decrease the permeability of the aquifer. In addition, an evalu-
ation of the compatibility of the TEA chosen with the site's
geochemistry should indicate whether undesired reactions can
occur.
The laboratory assessment for the removal of the parent
compound can measure the disappearance of the compound,
the rate of removal, and the production of daughter products.
The rate of removed usually reflects the laboratory conditions,
however, and cannot be extrapolated directly to the rate of
removal that would be expected in the field. Disappearance of
the parent compound may, by itself, not always indicate that
mineralization has occurred.
13.3.2 Compounds Appropriate to Consider for
Bioremediation
During the initial evaluations for bioremediation of a site,
existing information should be considered. Information about
the volubility of the compound to be degraded indicates the
potential availability of the compound to the microorganisms.
Previous evaluations of the biodegradation of the compound
often can be found in the scientific literature. These studies
can provide information about the inherent degradability of
the compound as well as the potential products of degradation.
Information about the environmental factors that upon stimu-
lation were critical to degradation also may be available.
Dragun (1988), for example, contains a list of organic com-
pounds and provides information about the conditions used in
the evaluations to develop the rates of biodegradation pre-
sented.
In general, hydrocarbons are good candidates for biore-
mediation. The review paper by Atlas (1981) and the books
edited by Gibson (1984) and Atlas (1984) provide an over-
view of the microbiological degradation of petroleum hydro-
carbons. Many components of fuel hydrocarbons, such as
benzene, toluene, and xylenes are degraded by a variety of
198
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microorganisms. Creosote, which is a by-product of the pro-
duction of coke from coal, is composed of a number of high
molecular weight hydrocarbons referred to as polycyclic aro-
matic hydrocarbons (PAHs). This complex mixture of hydro-
carbons has components that are biodegradable with the
degradation rate decreasing as the molecular weight of the
hydrocarbons increases. Generally, the PAHs with three rings
or less degrade at a greater rate than do the more complex
PAHS.
A variety of organic compounds can biodegrade in the
subsurface if the environmental conditions are appropriate.
For example, alcohols, glycols, ketones, phenols, chlorinated
phenols, and other organic compounds have the potential to
biodegrade. Some factors that may enhance biodegradation
are the water volubility and molecular weight of the com-
pounds. Increasing volubility enhances the potential for bio-
degradation assuming the concentration does not reach levels
toxic to the microorganisms. Increasing molecular weight or
branching of organic compounds may tend to slow the rate of
degradation.
Halogenated compounds generally tend to persist in aero-
bic environments, but continued research is providing evi-
dence that biological alternatives to these compounds may
exist. Under anaerobic conditions, several chlorinated com-
pounds have been shown to undergo transformation. For
example, tetrachloroethene (PCE) has been shown to be de-
chlorinated under environmental conditions that support the
growth of anaerobic bacteria. This process is known as reduc-
tive dechlorination and is given as follows:
PCE —> TCE + Cl —> DCE + Cl —> CE + Cl
—>C02+ Cl
The compounds produced are trichloroethene (TCE), the iso-
mers of dichloroethene (DCE), and chloroethene (CE). The
removal of the chlorine atoms enhances the potential for
aerobic microorganisms to degrade the daughter products.
DCE has a greater potential for aerobic degradation than does
PCE.
A method to enhance the aerobic degradation of DCE is
to create an environment for the growth of methane-utilizing
bacteria. The addition of methane to soils and aquifers typi-
cally results in the growth of these bacteria within several
days to a few weeks. These bacteria have been shown to
degrade a variety of halogenated compounds including TCE,
cis-DCE, tram-DCE, chloroform, dichloromethane, and 1,2-
dichloroethane (Henson et al, 1989). It seems plausible that
the series of reaction processes that enhances anaerobic reduc-
tive dechlorination of highly chlorinated compounds and yields
the less chlorinated compounds that can undergo aerobic
degradation may be a good mechanism to remove compounds
from the subsurface environment. The value for this treatment
process is further enhanced when the increased sorptive ca-
pacity of the higher chlorinated compounds is considered.
Utilizing the microorganisms in an in situ treatment process
can significantly expedite remediation.
Bioremediation of other halogenated compounds such as
polychlorinated biphenyls (PCBs) also can be considered. The
reductive dechlorination of PCBS was detected in the environ-
ment (Brown et al., 1987) and confirmed in the laboratory
(Quensen et al., 1988). The anaerobic reductive dechlorina-
tion process removes chlorines from the PCBS, thus reducing
potential toxicity and enhancing the aerobic degradability of
the compounds. Anaerobic biological treatment followed by
aerobic biological treatment is a technology that could remove
these chlorinated compounds from the environment in a cost-
effective and environmentally acceptable manner.
Bioremediation in the subsurface can remove a variety of
organic compounds. The evaluation of the bioremediation
process should include observation of the removal of the
organic compound(s) in a manner so as to provide a mass
balance. In the laboratory, mass balances can be approximated
with the use of proper abiotic controls. The use of abiotic
controls in the laboratory evaluation cannot be overempha-
sized. In the field, a mass balance can be approximated with
the collection of samples prior to remediation to evaluate the
amount of organic compound present. Samples collected sub-
sequent to the initiation of bioremediation can be evaluated
relative to the initial concentrations. If the bioremediation
effort is succeeding, a reduction in the concentration of the
organic compound should be observed. In areas not undergo-
ing bioremediation, the concentration of the organic com-
pound should remain relatively unchanged. If TEAs are added,
removal of these compounds also suggests biological activity.
The presence of metabolic intermediates also indicates that
biological processes are occurring. Other observations, such
as adaptation or acclimation or an increase in microbial activ-
ity of the compound being degraded, are positive indicators of
the enhancement of naturally occurring bacteria to achieve
bioremediation.
13.4 References
Alexander, M. 1981. Biodegradation of Chemicals of Envi-
ronmental Concern. Science 211:132-138.
Atlas, R.M. 1981. Microbial Degradation of Petroleum Hy-
drocarbons: an Environmental Perspective. Appl. Environ.
Microbiol. Vol. 45, pp. 180-209.
Atlas, R.M. 1984. Petroleum Microbiology. Macmillan, New
York.
Back, W. 1989. Early Concepts of the Role of Microorgan-
isms in Hydrogeology. Ground Water 27:618-622.
Balkwill, D.L. and W.C. Ghiorse. 1985. Characterization of
Subsurface Bacteria Associated with Two Shallow Aqui-
fers in Oklahoma. Appl. Environ. Microbiol. 50:560-588.
Beeman, R.E. and J.M. Suflita. 1987. Microbial Ecology of a
Shallow Unconfined Ground-water Aquifer Polluted by
Municipal Landfill Leachatc. Microb. Ecol. 14:39-54.
Brown, J.F., R.E. Wagner, H. Feng, DL. Bedard, M.J. Brennen,
J.C. Carnahan, and R.J. May. 1987. Environmental De-
chlorination of PCBS. Environ. Toxicol. Chem. 6579-
593,
199
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Caskey, W.H. and J.M. Tiedje. 1980. The Reduction of Ni-
trate to Ammonium by a Clostridium sp. Isolated from
Soil. J. Gen. Microbiol. 119:217-223.
Cline, P.V., J.J. Delfino, and T. Potter. 1988. Degradation and
Advection of 1,1,1-Tricholorethane in the Saturated Zone
Containing Residual Solvent. In: Superfund '88, Hazard-
ous Materials Control Research Institute, Silver Spring,
MD.
Dragun, J. 1988. The Soil Chemistry of Hazardous Materials.
Hazardous Materials Control Research Institute, Silver
Spring, MD.
Ehrlich, G.G., E.M. Godsy, D.F. Goerlitz, and M.F. Hull
1983. Microbial Ecology of a Creosote-Contaminated
Aquifer at St, Louis Park, Minnesota. Dev. Ind. Microbiol.
24:235-245.
Ehrlich, G.G., R.A. Schroeder, and P. Martin. 1985. Microbial
Populations in a Jet-Fuel Contaminated Shallow Aquifer
at Tustin, California. U.S. Geological Survey Open-File
Report 85-335.
Genthner, B.R.S., W.A. Price, and H.P. Pritchard. 1989.
Anaerobic Degradation of Chloroaromatic Compounds
in Aquifer Sediments under a Variety of Enrichment
Conditions. Appl. Environ. Microbiol. 55:1466-1471.
Ghiorse, W.C. and D.L. Balkwill. 1983. Enumeration and
Morphological Characterization of Bacteria Indigenous
to Subsurface Environments. Dev. Ind. Microbiol. 24:213-
224.
Ghiorse, W.C. and D.L. Balkwill. 1985. Microbiological Char-
acterization of Subsurface Environments. In: Ground-
Water Quality, C.H. Ward, W. Giger, and P.L. McCarty
(eds.), John Wiley & Sons, New York, pp. 536-556.
Ghiorse, W.C. and J.T. Wilson. 1988. Microbial Ecology of
the Terrestrial Subsurface. Adv. Appl. Microbiol. 33:107-
172.
Gibson, D. T. (ed.). 1984. Microbial Degradation of Organic
Compounds. Marcel Dekker, New York.
Grbic-Galic, D. and T.E. Vogel. 1987. Transformation of
Toluene and Benzene by Mixed Methanogenic Cultures.
Appl. Environ. Microbiol, 53:254-260.
Harvey, R.W., R.L. Smith, and L. George. 1984. Effect of
Organic Contaminants upon Microbial Distribution and
Heterotrophic Uptake in a Cape Cod, Massachusetts Aqui-
fer. Appl. Environ. Microbiol, 48:1197-1202.
Henson, J.M., M.V. Yates, and J.W. Cochran. 1989. Metabo-
lism of Chlorinated Methanes, Ethanes, and Ethylenes by
a Mixed Bacterial Culture Growing on Methane. J.
Indust. Microbiol. 4:29-35.
Humenick, M.J.H., L.N. Bitton, and C.F. Maddox. 1982.
Natural Restoration of Ground Water in UCG. In Situ
6:107-125
Jamison, V.W., R.L. Raymond, and J.O. Hudson. 1975. Bio-
degradation of High-Octane Gasoline in Groundwater.
Dev. Ind. Microbiol. 16:305-312.
Knowles, R. 1982. Denitrification. Microbiol. Rev. 46:43-70.
Ladd, T.I., et al. 1982. Heterotrophic Activity and Biodegra-
dation of Labile and Refractory Compounds by Ground
Water and Stream Microbial Populations. Appl. Environ.
Microbiol. 44:321-329.
Lee, M.D. and C.H. Ward. 1985. Biological Methods for the
Restoration of Contaminated Aquifers. Environ. Toxicol.
Chem. 4:721-726.
Lee, M.D., J.M. Thomas, R.C. Borden, P.B. Bedient, J.T.
Wilson, and C.H. Ward. 1988. Biorestoration of Aquifers
Contaminated with Organic Compounds. CRC Crit. Rev.
Environ. Control 18:29-89.
McCarty, P.L. 1975. Energetics of Organic Matter Degrada-
tion. In: Water Pollution Microbiology, Vol. 1, R. Mitchell
(ed.), Wiley-Interscience, New York, pp. 91-110.
McGinnis, G. D., H. Borazjani, L.K. McFarland, D.F. Pope,
and D.A. Strobcl. 1988. Characterization and Laboratory
Soil Treatability Studies for Creosote and Pentachloro-
phenol Sludges and Contaminated Soil. EPA/600/2-88-
055 (NTIS PB89-109920).
McNabb, J.F. and G.E. Mallard. 1984. Microbiological Sam-
pling in the Assessment of Groundwater Pollution. In
Groundwater Pollution Microbiology, G. Bitton and C. P.
Gerba (eds.), John Wiley & Sons, New York, pp. 235-
260.
Postgate, J.R. 1979. The Sulphate-Reducing Bacteria. Cam-
bridge University Press, Cambridge.
Quensen, J.F., J.M. Tiedje, and S.A. Boyd. 1988. Reductive
Dechlorination of Polychlorinated Biphenyls by Anaero-
bic Microorganisms from Sediments. Science 242:752-
754.
Raymond, R. L., V.W. Jamison, and J.O. Hudson. 1976. Ben-
eficial Stimulation of Bacterial Activity in Ground Water
Containing Petroleum Products. AICE Symposium Se-
ries 73:390-404.
Roberts, P. V., L. Semprini, G.D. Hopkins, D. Grbic-Galic,
P.L. McCarty, and M. Reinhard. 1989. In-Situ Restora-
tion of Chlorinated Aliphatics by Methanotrophic Bacte-
ria. EPA/600/2-89-033 (NTIS PB89-219992).
Smolenski, W.J. and J.M. Suflita. 1987. Biodegradation of
Cresol Isomers in Anoxic Aquifers. Appl. Environ.
Microbiol. 53:710-716.
200
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Suflita, J.M., S.A. Gibson, and R.E. Beeman. 1988. Anaero-
bic Biotransformation of Pollutant Chemicals in Aqui-
fers. J. Ind. Microbiol. 3:179-194.
Thomas, J.M. and C.H. Ward. 1989. In Situ Biorestoration of
Organic Contaminants in the Subsurface. Environ. Sci.
Technol. 23:760-766.
Updegraff, D.M. 1982. Plugging and Penetration of Petro-
leum Reservoir Rock by Microorganisms. In: Proc. 1982
Int. Conf. on Microbial Enhancement of Oil Recovery.
Ventullo, R.M. and R.J. Larson, 1985. Metabolic Diversity
and Activity of Heterotrophic Bacteria in Ground Water,
In Environ. Toxicol. Chem. 4:759-771.
Vogel, I.E. and D. Grbic-Galic. 1986. Incorporation of Oxy-
gen into Toluene and Benzene During Anaerobic Fer-
mentative Transformation. Appl. Environ. Microbiol.
52:200-202.
Waksman, S.A. 1916. Bacterial Numbers in Soil, at Different
Depths, and in Different Seasons of the Year. Soil Sci-
ence 1:363-380.
Webster, J.J., G.J. Hampton, J.T. Wilson, W.C. Ghiorse, and
F.R. Leach. 1985. Determination of Microbial Numbers
in Subsurface Environments. Ground Water 23:17-25.
Wilson, B.H. and J.F. Rees. 1985. Biotransformation of Gaso-
line Hydrocarbons in Methanogenic Aquifer Material. In:
Proc. NWWA/API Conf. on Petroleum Hydrocarbons
and Organic Chemicals in Ground Water-Prevention,
Detection and Restoration, National Water Well Associa-
tion, Dublin, OH, pp. 128-139.
Wilson, J.T. and C.H. Ward. 1987. Opportunities for Biorec-
lamation of Aquifers Contaminated with Petroleum Hy-
drocarbons. Dev. Indust. Microbiol. 27:109-116.
Wilson, J.T., J.F. McNabb, D.L. Balkwill, and W.C. Ghiorse.
1983. Enumeration and Characterization of Bacteria In-
digenous to a Shallow Water-Table Aquifer. Ground
Water 21:134-142.
Wilson, J.T. J.F. McNabb, J.W. Cochran, T.H. Wang, M.B.
Tom son, and P.B Bedient. 1985a. Influence of Microbial
Adaptation on the Fate of Organic Pollutants in Ground
Water. Environ. Toxicol. Chem. 4:721-726.
Wilson, J.T., M.J. Noonan, and J.F. McNabb. 1985b. Biodeg-
radation of Contaminants in the Subsurface. In: Ground-
Water Quality, C.H. Ward, W. Giger, and P.L. McCarty,
(eds.), John Wiley & Sons, New York, pp. 483-498.
Wilson, J.T., L.E. Leach, M. Henson, and J.N. Jones. 1986. In
Situ Biorestoration as a Ground Water Remediation Tech-
nique. Ground Water Monitoring Review 6(4):56-64.
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PART III: SOIL AND GROUND-WATER REMEDIATION
Chapter 14
Soil and Ground-Water Remediation: Basic Approaches
Ronald C. Sims and Judith L. Sims
Subsurface remediation includes identifying, quantify-
ing, and controlling contaminant source(s); considering cleanup
levels required for each medium (air, soil, and ground water)
to protect human health and the environment; and selecting
treatment technologies based on information obtained con-
cerning source(s) and cleanup levels. The challenge is to
effectively relate site characterization activities to selecting
the most appropriate remediation technologies for contami-
nated soils and ground water at hazardous waste sites. Effec-
tively relating these activities with technology selection
improves the efficiency, purpose, and results of both site
characterization and remediation technique selection. This
chapter addresses specific subsurface physical, chemical, and
biological processes that have been discussed in previous
chapters within the context of(l) site characterization require-
ments, (2) evaluation and selection of remediation techniques
and treatment trains utilizing several techniques, and (3) de-
sign of monitoring programs.
There is currently a lack of methods and approaches for
evaluating and selecting remedial technologies for site-spe-
cific scenarios in the area of subsurface remediation, includ-
ing soil and ground-water remediation. This chapter presents
a rational approach for addressing soil and ground-water
remedial technologies, including evaluating and selecting new
technologies as they become available to the user community.
Specific soil and aquifer remediation techniques, including
applications and limitations, also are discussed.
14.1 Conceptual Approach to Soil and Ground-
Water Remediation
A conceptual framework for soil remediation technique
evaluation, selection, and monitoring, based on current infor-
mation and activities employed at hazardous waste sites is
proposed. The conceptual framework is the chemical mass
balance, the cornerstone of science and engineering research
and industry. The concept of a chemical mass balance is
familiar to professionals trained in the physical or life sci-
ences or in engineering. It provides a rational and fundamental
basis for asking specific questions and obtaining specific
information that is necessary for determining fate and behav-
ior, for evaluating and selecting treatment options, and for
monitoring treatment effectiveness at both laboratory-scale
and field-scale. A mass balance approach also meets the goal
of obtaining quantitative accuracy about the amount of con-
taminants initially present at an uncontrolled site. While a
mass balance, or materials balance, is routinely conducted on
aboveground treatment processes (Bailey and Ollis, 1986;
Benefield et al, 1982; Corbitt, 1989; Metcalf and Eddy, Inc.,
1979), and for ground-water processes (Willis and Yeh, 1987;
Wilson et al., 1989), a mass balance approach has generally
not been applied to the soil environment or to the subsurface/
surface system to link characterization activities and treat-
ment technology selection. The information needed to con-
struct a mass balance for contamination at a site simultaneously
addresses site characterization and remediation evaluation
and selection.
The conceptual approach for the soil and ground-water
subsurface environment at a contaminated site is illustrated in
Figure 14-1. The contaminated subsurface is a system gener-
ally consisting of two phases (solid and fluid) and five com-
partments (gas, an inorganic mineral solid compartment an
organic matter solid compartment, water, and oil [NAPL])
(Sims et al., 1989). Generally NAPLs are subdivided into two
classes: those that are lighter than water (LNAPLs), and those
with a density greater than water (DNAPLs). LNAPLs in-
clude hydrocarbon fuels, such as gasoline, heating oil, kero-
sene, jet fuel, and aviation gas. DNAPLs include chlorinated
hydrocarbons, such as 1,1,1 -trichloroethane, carbon tetrachlo-
ride, chlorophenols, chlorobenzenes, tetrachloroethylene, and
polychlorinated biphenyls (PCBs).
Specific subsurface processes concerning water move-
ment, sampling, sorption and reaction, and degradation are
discussed in the previous chapters. The processes and termi-
nology described in the previous chapters will be used in this
chapter for the discussion of the components of a mass
balance and the mass balance approach to evaluation and
selection of soil remediation techniques.
Interphase transfer potential for waste constituents among
oil (waste or NAPL), water, air, and solid (organic and inor-
ganic) phases of a subsurface system is affected by the relative
affinity of waste constituents for each phase shown in Figure
14-1, and may be quantified through calculation of distribu-
tion coefficients (Loehr, 1989; Sims et al., 1988; U.S. EPA,
203
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Organic
Water
Fluid Phase
Gas
Solid Phase
Inorganic
NAPL
Figure 14-1. Mass balance conceptual framework for the soil end ground-water subsurface environment at a contaminated site.
1986). Distribution coefficients are calculated as the ratio of
the concentration of a chemical in the soil (or aquifer materi-
als), oil, or gas phases to the concentration of a chemical in the
water phase. A waste chemical, depending on its tendency to
be associated with each phase, will distribute itself among the
phases, and can be quantified in terms of distribution coeffi-
cients. Distribution coefficients are available for a variety of
chemicals and can be expressed as ratios of the concentrations
of a chemical between two phases in the subsurface
Kd= Concentration in solid phase/Concentration in aque-
ous phase
K0= Concentration in oil phase/Concentration in aqueous
phase
Kh= Concentration in air phase/Concentration in aqueous
When distribution coefficients are not available, they can be
estimated using structure-activity relationships (SARs) or can
be determined in laboratory tests (Sims et al., 1988). For
additional detail concerning these processes, see Chapters 10
and 11.
Distribution coefficients have been used most success-
fully with organic chemicals. However, since metals distrib-
ute among the phases of the subsurface systems described
previously, distribution coefficients also may be used, along
with multiphase metal speciation information (Sims et al.,
1984), to evaluate metal distribution in a contaminated sub-
surface system. For additional detail concerning these pro-
cesses see Chapter 12.
Knowledge of migration and distribution of chemicals
and chemical intermediates among the phases and compart-
ments of a contaminated subsurface system (illustrated in
Figure 14-2) provides fundamental information about the fate
and behavior of contaminants, which can be used for selecting
and evaluating subsurface remedial techniques. Retardation
of the downward transport (leaching potential) and upward
transport (volatilization potential) is referred to as immobili-
zation of waste constituents, and has been related to the
subsurface organic matter content, especially for hydrophobic
chemicals (Nkedi-Kizza et al., 1983), soil moisture (Mahmood,
1989), and presence and concentration of organic solvents
(Mahmood and Sims, 1986; Rao et al., 1985).
In summary, subsurface processes described above, com-
bined with information about the movement of fluids as
discussed in Chapters 4, 5, and 6 (gases, aqueous phase, and
pure product flow) in the unsaturated and saturated zones,
provide the inputs into the chemical mass balance that can be
used for (1) characterizing a site; (2) assessing the problem of
mobility; (3) evaluating treatment techniques; and (4) identi-
204
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Volatilization
Hazardous
Contaminant
t\ t
Mineralization
Biomass
Soil Interactions
Phases: Solid Liquid Gas
Intermediate
Products
Leaching
Figure 14-2. Interphase transfer potential of chemicals in the
subsurface (from Sims et al., 1990).
fying chemicals in specific phases for monitoring treatment
effectiveness.
14.2 Methodology
Using the chemical mass balance approach, the authors of
this Handbook developed a methodology for integrating data
collection activities at CERCLA sites to address simultaneous
site characterization and remediation technique selection, The
proposed methodology consists of four elements: (1) charac-
terization, (2) assessment of the problem, (3) treatment (train)
selection, and (4) monitoring treatment performance (Figure
14-3). The first element involves characterization in the con-
text of waste/subsurface/site interactions to address the ques-
tion, "Where is the contamination and in what form(s) does it
exist?" The second element, assessment of the problem, uti-
lizes subsurface fate and behavior information to address the
question "Where is the contamination going under the influ-
ence of natural processes?" The problem can be define in the
context of mobility versus degradation for chemicals at a site.
Using mathematical models or other tools, the chemicals can
be ranked in order of their relative tendencies to leach, to
volatilize, to move in a NAPL phase and to remain in-place
under site-specific conditions. Containment and/or treatment
options then can be selected that are chemical-specific and
that address specific escape and attenuation pathways (third
Methodology for Integrating Site Characterization with Subsurface Remediation
Characterization
Site
i
So/7
1
Problem Assessment
Treatment (train)
Monitoring
Figure 14-3. Methodology using mass balance approach for integrating data collection activities at a contaminated site.
205
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element). Therefore, treatment trains can be selected to ad-
dress specific waste phases at specific times during remedia-
tion (volatile, leachate, solid phase, and pure product), with
the selection based upon results of a mass balance evaluation
through time to identify the fate of each waste phase. Finally
monitoring programs can be designed for specific chemicals
in specific phases in the subsurface at specific times (fourth
element).
The approach for using the methodology described above
consists of applying a mass balance for each element of the
methodology. This approach assists in the collection of spe-
cific information that is transferable among all four elements
of the methodology, and also addresses the technical issues of
soil remediation within the context of regulatory goals.
14.2.1 Site Characterization
Identifying waste sources by subsurface phases, i.e., iden-
tification and amount (if possible) of waste constituents asso-
ciated with solid and fluid phases (Figure 14-1), allows
assessment of the magnitude (mass) and physical form(s) of
waste that must be treated. This assessment comprises the first
step in the mass balance characterization of waste sources at a
site.
Wastewater historically has been characterized and sub-
sequently treated in terms of its interaction and potential
impact of the assimilative capacity of surface water receiver
systems, generally rivers or lakes (e.g., requiring measure-
ment of characteristics such as oxygen-demanding substances,
nutrients, and levels of substances toxic to aquatic organisms).
However, a waste characterization program at a hazardous
waste site addresses the vadose zone and ground water, in
addition to surface water, as the receiver systems (e.g., requir-
ing measurement of characteristics that reflect individual
chemical mobility and destruction in the subsurface environ-
ment and those that affect human health as well as character-
istics that affect environmental toxicity). Also, it describes the
behavioral interaction of waste chemicals in each surface and
subsurface phase. Thus, hazardous waste is more appropri-
ately characterized in terms of the interaction and potential
impact on the subsurface assimilative capacity.
Specific site characteristics important for describing and
assessing the environmental behavior and fate of organic
constituents in the soil and subsurface are listed in Table 14-1.
For each chemical, or chemical class, required information
includes (1) characteristics related to potential leaching (e.g.,
water solubility, octanol/water partition coefficient, solid sorp-
tion coefficient); (2) characteristics related to potential vola-
tilization (e.g., vapor pressure, relative volatilization index);
(3) characteristics related to potential degradation (e.g., half-
life, degradation rate, degradability index); and (4) character-
istics related to chemical reactivity (e.g., hydrolysis half-life,
soil redox potential) (Sims et al., 1984). The information
presented in Table 14-1 also is used to assess problem(s)
concerning migration potential at a site and to evaluate and
select containment- and treatment-management options.
If the distribution of waste chemicals among phases that
comprise the soil and subsurface at a site are determined, then
potential pathways of transport, or escape, from a site can be
indicated. Therefore, exposure pathways for human health
and the environment may be evaluated, i.e., risk assessment
can be made. Through a determination of subsurface flow
conditions as part of site characterization activities (aqueous,
gas, and pure product flow in the vadose zone and aqueous
plume and pure product movement in the saturated zone), the
mass of material moving through a site and potential move-
ment off site can be assessed:
concentration (mass/vol) X rate of flow (vol/time)
= mass flow at site (mass/time)
This information is combined with additional information,
discussed in the next section, that is needed to assess the
problem(s) with respect to treatment technique selection.
The U.S. Environmental Protection Agency (EPA's) Rob-
ert S. Kerr Environmental Research Laboratory, as part of its
Superfund Technology Support Center Program activities,
provides assistance to EPA regional offices and state regula-
tory agencies about appropriate site characterization activities
at Superfund sites and other uncontrolled hazardous waste
sites to support selection of effective remediation tcchnol-
gies. Table 14-2 presents examples of recommended site
evaluation and characterization actions as related to the use of
soil and the subsurface as the receiver system at uncontrolled
hazardous waste sites (Scalf and Draper, 1989).
14.2.2 Assessment of Problem
Assessment of the contamination involves organizing the
information obtained from site characterization activities to
evaluate the transport and degradation behavior of each chemi-
cal of concern at a site under consideration. Specifically, the
rate of transport can be compared with the rate of degradation
to determine if transport is significant relative to degradation.
This approach to problem(s) assessment will allow chemicals
to be prioritized individually according to (1) magnitude and
rate of transport (escape) from a site, (2) persistence, and (3)
pathway(s) of migration from a site. Treatment technique
evaluation and selection then can be based upon specific
combinations of chemical and physical phase-migration
pathway.
Interfacing subsurface-based behavioral characteristics
of specific contaminants (Table 14-1) with specific site and
subsurface properties allows an assessment of the problem(s)
related to contamination of other media (due to mobility),
including the ground water under the contaminated area, the
atmosphere over the site or at the site boundaries, surface
waters, and/or persistence of chemicals at a site. Pathways of
movement and potential mechanisms of removal of contami-
nants at a specific site are illustrated in Figure 14-2. This
element of the methodology functions to identify chemicals
that will (1) migrate upward (volatilization), (2) migrate down-
ward (leaching), (3) migrate laterally (aqueous plume and
pure product), (4) degrade, and (5) remain at the site as
persistent chemicals. By ranking the chemicals in the order in
which Ihey migrate or persist, chemicals can be prioritized
with regard to urgency for treatment and for monitoring.
206
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Table 14-1. Subsurface-Based Waste Characterization
Chemical Class
Acid
Base
Polar neutral
Nonpolar neutral
Inorganic
Chemical Properties
Molecular weight
Melting point
Specific gravity
Structure
Water solubility
Chemical Reactivity
Oxidation
Reduction
Hydrolysis
Precipitation
Polymerization
Soil Sorpiion Parameters
Freundlich sorption constants (K, N)
Sorption based on organic carbon content (K^
Octanol water partition coefficient (KaJ
Soil Degradation Parameters
Half-life (tia)
Rate constant (first order)
Relative biodegradability
Soil Volatilization Parameters
Air: water partition coefficient (KJ
Vapor pressure
Henry's Law constant
Sorption based on organic carbon content (K^)
Water solubility
Soil Contamination Parameters
Concentration in soil
Depth of contamination
Date of contamination
Source: Sims et al., 7984
Waste characteristics identified in Table 14-1, including
potential sorption, degradation, and volatilization at a site, can
be determined in laboratory mass balance tests, using waste/
soil mixtures from a site. These characteristics can be used to
evaluate the fate of the waste at the site, and to generate
specific data that can be used to develop treatment approaches.
Figure 14-4 illustrates a laboratory flask apparatus that can be
used to develop a chemical mass balance by measuring inter-
phase transfer potential of chemicals as well as degradation
potential at a site (Park et al, 1990).
The contaminated material is placed in a flask, which is
then closed and incubated under controlled conditions for a
period of time. During the incubation period, air is drown
through the flask and then through a sorbent material. Volatil-
ized materials are collected by the sorbent and are measured to
estimate volatilization loss of the constituents of interest. At
the end of the incubation period, a portion of the contaminated
soil is treated with an extracting solution to determine the
extent of loss of the constituents in the soil matrix. This loss
can be attributed to degradation and possible immobilization
in the soil materials. It is necessary to select an appropriate
extracting solution and procedure to maximize constituent
recovery from a soil-waste mixture (Coover et al., 1987).
Another portion of the soil is leached with water to determine
leaching potential of the remaining constituents. Abiotic and
biological processes involved in removal of the parent com-
pound are evaluated by comparing microbially active soil/
waste mixtures with mixtures that have been treated with a
microbial poison, e.g., mercuric chloride or propylene oxide.
Samples generated from the different phases of the system in
microcosm mass balance studies identified above can be
analyzed for intermediate degradation products and used in
bioassay studies to provide information concerning transfor-
mation and detoxification processes.
The use of a procedure incorporating features illustrated
by the use of this microcosm (Figure 14-4) is crucial to obtain
a materials balance of waste constituents in the subsurface
system, Examples of such protocols may be found in EPA
guidance documents and research reports (Loehr, 1989; Sims
et al., 1988; EPA, 1986; and Park et al., 1990). Contaminated
materials also can be spiked with radiolabeled chemicals;
tracking the fate of the chemicals as they move through the
multiple phases of the soil system also provides a materials
mass balance.
The mass balance approach identified above usually rep-
resents optimum conditions with respect to mixing, contact of
sol id materials with waste constituents, and homogeneous
conditions throughout the laboratory microcosm; therefore, it
does not incorporate site nonhomogeneity in the evaluation.
This aspect must be defined during site characterization ac-
tivities and evaluated with regard to potential effect on fate
and behavior regarding migration and persistence at the site
(problem assessment).
In addition to the laboratory tests described, bench-scale
reactors, pilot-scale reactors and/or field-scale plots may be
used to generate mass balance information for problem as-
sessment. The set of experimental conditions (e.g., tempera-
ture, moisture, waste concentration) under which the studies
were conducted and experimental results should be presented.
Information from the performance of site characterization
and experimental mass balance studies may be integrated with
the use of comprehensive mathematical modeling to aid in
problem assessment. In general, models are used to analyze
the behavior of an environmental system under both current
(or past) and anticipated (or future) conditions (Donagian and
Rao, 1986). A mathematical model provides a tool for (1)
integrating degradation and partitioning processes with site-,
soil-, and waste-specific characterization; (2) simulating the
behavior of waste constituents in a contaminated soil; and (3)
predicting the pathways of migration through the contami-
nated area, and therefore pathways of exposure to humans and
to the environment. DiGiulio and Suffet (1988) and Weaver et
al. (1989) have presented guidance on the selection of appro-
priate subsurface zone models for site-specific applications,
focusing on recognition of limitations of process descriptions
of models and difficulties in obtaining input parameters re-
quired by these process descriptions.
The Regulatory and Investigative Treatment Zone Model
(RITZ), developed at the EPA's Robert S. Kerr Environmen-
tal Research Laboratory by Short (1986) is an example of a
vadose zone model that has been used to describe the potential
fate and behavior of organic constituents in a contaminated
soil system (U.S. EPA, 1988a). The RITZ Model is based on
207
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Table 14-2. Examples of Suggested Site Characterization Activities Based on Soils and Subsurface Materials as Waste-Receiver
Systems
Site
USEPA
Region
Contaminants
Recommended Site Evaluation
and Characterization Actions
Stamina Mils Superfund Site
North Smithfield, Rl
W.R. Grace & Co. Superfund Site
Acton, MA
Somersworth Landfill
Somersworth, NH
Nascolite Superfund Site
Millsville and Vineland, NJ
Drake Chemical Superfund Site
Lock Haven, PA
Tyson's Dump
Anderson Development Co.
Spill Cleanup
Adrian, Ml
Montrose Chemical Site
Los Angeles, CA
Time Oil Site
Tacoma, WA
Frontier Hard Chrome
Vancouver, WA
TCE
Acetone, benzene,
toluene
Arsenic and organic
compounds
Methyl methacrylate
(MMA)
III Various wastes
III 1,2,3-trichloro-propane,
xylene, toluene, and
ethyl-benzene
V 4,4-methylene-bis-2-
chloroaniline (MBOCA)
IX DDT
PCE, PCA, & TCE
Chromium, lead,
nickel, S cyanide
Determination of soil-water partition coefficients;
investigation of soil physical and hydraulic properties;
Simulation of contaminant transport
Selection of soil physical properties forfugacity
modeling
Selection of leaching test suitable for high organic matter
content soils to provide data for estimation of migration
potential
Evaluation of residual soil concentrations during
groundwater fluctuations; Development of appropriate
extraction technologies based on chemical properties of
MMA
Development of laboratory procedures: Determination
of site-specific partition coefficients; Development of
ContPro Model (revised version of RITZ Model)
Determination of causes for plugging of SVE extraction
wells with tarry materials
Recommendation of use of site-specific biotreatability
study to determine feasibility of use of soil bioremediation
Recommendation to design of laboratory soil biotreatment
feasibility studies
Development of soil-water and soil-air partitioning
relationships for implementation of SVE
Development of estimates ofleachate concentrations of
contaminants at equilibrium between soil and water
Source: Scalf and Draper, 1989
Influent
Purge Gas
••Effluent Purge Gas
Sorbent
Tubes
Constant
Flow
Sample
Pump
, Soil/Waste
Mixture
Effluent Purge Gas
Figure 14-4. Laboratory flask apparatus used for mass balance measurements (from Park et al., 1990).
208
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an approach developed by Jury et al. (1983). An expanded
version of RITZ, the Vadose Zone Interactive Processes (VIP)
model, incorporates predictive capabilities for the dynamic
behavior of organic constituents in unsaturated soil systems
under conditions of variable precipitation, temperature, and
waste concentrations (McLean et al., 1988; Stevens et al.
1988, 1989; Symons et al., 1988; U.S. EPA, 1986). Both the
RITZ and VIP models simulate vadose zone processes, in-
cluding volatilization, degradation, sorption/desorption, ad-
vection, and dispersion (Grenney et al., 1987).
For example, the VIP model was used to evaluate the
relative tendencies for a group of pesticides to volatilize and
to leach under specific waste-soil conditions (McLean et al.,
1988). Information input into the model included half-life
(measured in laboratory tests), distribution coefficients
(Kd,Kh,Ko) (calculated), soil texture and moisture (measured),
and site-specific climatic data (rainfall and temperature). Re-
sults are presented in Table 14-3. The ranking of pesticides
provided by the model indicated that the tendency of the
pesticides to volatilize was not similar to their tendency to
leach (McLean et al. 1988). This information can be used to
assess which chemicals are likely to volatilize first, which
chemicals are likely to leach first, and which chemicals are
persistent under site-specific conditions. In addition to assist-
ing in the problem assessment step of the methodology,
mathematical models also can be used to design studies for
evaluation and selection of treatment options for these chemi-
cals, as well as to design monitoring strategies (i.e., which
chemicals to monitor in which media).
With regard to ground-water models that can be used as
part of the problem assessment, the International Ground
Water Modeling Center (IGWMC), through EPA, published
information about the kinds and availability of models, their
specific characteristics, and the information, data, and techni-
cal expertise needed for their operation (U.S. EPA, 1988 b).
Ground-water models also have been addressed within the
context of scientific and regulatory applications, with selected
case studies, by the National Research Council (NAS, 1990).
Table 14-3.
Ratios of Concentration of Pesticides Between
Water/Soil and Air/Soil at 15 cm After 81 Days
(Ranked in Order from Greatest Potential for
Leaching and Volatilization to Least Potential)
Leaching potential
(concentration in soil
water/concentration
Volatilization
potential
(concentration in soil
air/concentration
Pesticide
Disulfoton
Phorate
Methylparathion
Toxaphene
Endosulfan
Parathion
Heptachlor
Aldrin
in soil)
330
23
4.8
0.5
0.12
0.06
0.06
0.0009
Pesticide
Toxaphene
Disulfoton
Phorate
Heptachlor
Endosulfan
Aldrin
Methylparathion
Parathion
in soil)
7.4
3.6 x1O'
5.2X1&*
5.5 x 1O3
4.0x10*
2.0 x ?a5
1.2 x 10s
1.6 x 10"
McLean et al., 1988
A numerical model, BIOPLUME, was developed to simu-
late oxygen-limited biodegradation in ground-water environ-
ments. BIOPLUME simulates advection, dispersion, and
retardation processes as well as the reaction between oxygen
and the contaminants under steady, uniform flow (Rifai et al.,
1989). BIOPLUME was applied to an aviation gasoline spill
site at Traverse City, Michigan. Model predictions for the
rates of mass loss closely matched calculated rates from field
data.
14.2.3 Treatment Approaches
Information obtained from an integrated assessment (mod-
eling) of the problem (migration and persistence), based upon
a thorough characterization of waste/soil/site interactions, can
be used to select treatment approaches for further evaluation
with respect to technical and cost-effectiveness factors. Re-
sults of characterization and assessment efforts can aid in the
identification of constituents that will require treatment in the
following phases: (1) air (volatile) phase, (2) leachate phase,
and (3) solid (soil) phase. This approach allows evaluation
and comparison of different treatment systems identified pre-
viously (in situ and prepared bed). Specifically, if treatment is
required, the information is used to (1) determine containment
requirements to prevent contamination of offsite receiver
systems; (2) develop techniques to maximize mass transfer of
chemicals affecting a process (e.g., affecting microbial activ-
ity through addition of mineral nutrients, oxygen, additional
energy sources, pH control products, or removal of toxic
products in order to enhance bioremediation); and (3) design a
cost-effective and efficient monitoring program to evaluate
effectiveness of treatment.
Containment Requirements. If the major pathway of trans-
port is volatilization, containment and treatment to control
volatilization is required. An inflatable plastic dome erected
over a contaminated site is a containment method that has
been used to control escape of volatile constituents at hazard-
ous waste silts (St. John and Sikes, 1988). Volatiles are drawn
from the dome through a conduit and treated in an aboveground
treatment system. If leaching has been identified as an impor-
tant factor, control of soil water movement should be imple-
mented. For example, if contaminated materials are expected
to leach downward from the site, run-on and run-off controls
can be implemented, or the contaminated materials can be
temporarily removed from the site and a plastic or clay liner
can be placed under the site (Lynch and Genes, 1989; Ross et
al., 1988). When downward as well as upward migration are
significant, both volatilization and leaching containment sys-
tems can be installed. Some hydrophobic chemicals do not
tend to volatilize or to leach but are persistent within the soil
solid phase; therefore, containment efforts may not be re-
quired. With regard to the saturated zone, containment is
generally accomplished by physical barriers (e.g., slurry walls,
sheet pilings, grout curtains) or hydraulic barriers (e.g., pump
ing systems, french drains).
Maximizing Chemical Mass Transfer. An area of sig-
nificant research concerns delivery and recovery technologies
for maximizing mass transfer of chemicals that affect the rate
and/or extent of treatment. Murdoch et al. (1988) discussed
209
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delivery and recovery technologies, many of which are de-
rived from the petroleum and mining industries. While a
liquid phase is usually employed for delivery of chemicals,
some technologies utilize vapor and solid phases for delivery.
Principal recovery technologies involve hydraulic, thermal,
and chemical systems. Delivery and recovery techniques are
important in influencing the success of technologies, includ-
ing bioremediation, vapor extraction, and solidification/stabi-
lization. Specific delivery and recovery systems for in situ
treatment systems identified by EPA include hydraulic frac-
turing, radial well drilling, ultrasonic methods, kerf ing, jet-
induced slurry methods, carbon dioxide injection, hot brine
injection, and cyclic pumping (U.S. EPA, 1990).
14.2.4 Monitoring Program
A mass balance approach to monitoring, the fourth ele-
ment in the methodology (Figure 14-3), can be performed at
laboratory, pilot, and field scales. Monitoring efforts can be
focused on the appropriate environmental phase to evaluate
treatment effectiveness for specific chemicals. If a compre-
hensive and thorough evaluation of a specific contaminated
system has been conducted, not all chemicals may need to be
monitored in each phase. Specific chemicals wall be associ-
ated with specific phases; therefore, a monitoring plan can be
designed that is chemical/phase specific. This approach also
focuses analytical efforts so that methods of development are
chemical- and phase-specific.
The level of contamination associated with a particular
treatment technology requires monitoring. In addition, the
treatment system components, including delivery and recov-
ery systems, maintenance, and structures such as infiltration
galleries must be monitored.
14.3 Selection of Treatment Methods
14.3.1 Utility of Mathematical Models
A critical and cost-effective use of modeling in treatment
(train) selection and evaluation is for analysis of proposed or
alternative future conditions i.e., the model is used as a
management or decision-making tool to help answer "what if"
questions (Donagian and Rao, 1986). Models also may be
used to approximate and estimate the rates and extent of
treatment that may be expected at the field-scale under vary-
ing conditions. Attempting to answer such questions through
data collection programs would be expensive and practically
impossible in many situations. For example, information can
be generated to evaluate the effects of using different ap-
preaches for enhancing microbial activity and for accelerating
biodegradation and detoxification of the contaminated area by
altering environmental conditions that affect microbial activ-
ity. Therefore, modeling may be used to assist in the design of
treatability studies for considering and evaluating the applica-
tion of different treatment technologies, and therefore to assist
in focusing available resources (time and money). Section
14.2.2 (Assessment of Problem) provides more information
on the existence, applications, and limitations of mathemati-
cal models for vadose zone and ground-water analysis and
management.
14.3.2 Treatability Studies
Treatability studies can be used for evaluating and com-
paring rate and extent of remediation among several technolo-
gies and also to provide specific information about the potential
application of treatment technologies at field scale. Treatabil-
ity studies can be conducted in laboratory microcosms or
bench-scale reactors, pilot-sale facilities, or in the field. Labo-
ratory treatability studies are generally screening studies used
to (1) establish the validity of a technology, (2) generate data
that can be used as indicators of potential to meet performance
goals, and (3) identify parameters for investigation during
bench- or pilot-scale testing. Laboratory treatability studies
are generally not appropriate for generating design or cost
data (U.S. EPA, 1989). Pilot-scale testing is conducted to
generate information on quantitative performance, cost, and
design information. Three proposed categories of treatability
testing and associated descriptions are included in Table 14-4
(U.S. EPA, 1989b).
Treatability study results are commonly used to provide
information on rates and extent of treatment of hazardous
organic constituents when mass transfer rates of potential
limiting substances are not limiting the treatment. Treatability
studies also usually represent optimum conditions with re-
spect to mixing, contact of soil solid materials with waste
constituents and with microorganisms, and homogeneous con-
ditions throughout the microcosm. Therefore, treatability stud-
ies provide information concerning potential levels of
treatment. Rates and extent of remediation in a prepared bed
Table 14-4. General Comparison of Laboratory Screening, Bench-Scale Testing, and Pilot-Scale Testing
Tier
Laboratory screening
Bench-scale testing
Pilot-scale testing
Type
of data Critical
generated parameters
Qualitative
Quantitative
Quantitative
several
Few
Few
No. of
replicates
Single/
duplicate
Duplicate/
triplicate
Triplicate
or more
Study size
Jar tests or
beaker studies
Bench-top
(some larger)
Pilot-plant
(onsite or offsite)
Usual
process
type
Batch
Batch or
continuous
Batch or
continuous
Waste
stream
volume
Small
Medium
Large
lime
required
Hours/
day
Days/
week
Weeks/
month
Cost, $
10,000-
50,000
50,000-
250,000
250,000
1,000,000
Source: U.S. EPA, 1989
210
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or in situ system are generally limited by accessibility and rate
of mass transfer of chemical substances to the contaminated
soil and removal of inhibitory microbial degradation products
(Symons and Sims, 1988).
Information from mass balance treatability studies, in-
cluding laboratory screening-, bench- and pilot-scale studies,
is combined with information about site and waste character-
istics to determine applications and limitations of each tech-
nology. Information obtained from treatability studies should
be focused on identifying ultimate limitations to the use of a
remediation technology at a specific site. Limitations are
usually related to (1) time required for cleanup, (2) level of
cleanup attainable, and (3) cost of cleanup (Sims et al, 1989).
14.3.3 Treatment Trains
The use of treatment trains also is important to consider
in an engineering approach for using treatment techniques for
subsurface site remediation. For example, vacuum extraction
is known to be applicable to unsaturated sites characterized by
permeable materials containing volatile chemicals. Vacuum
extraction also can be used for the degradation of more
semivolatile chemicals. This degradation is accomplished by
providing a source of oxygen (air) to the subsurface environ-
ment microorganisms where anoxic conditions exist due to
relative slow replenishment of oxygen through atmospheric
diffusion. This is an example of the use of one technology for
the treatment of both volatile and semivolatile chemicals in
the subsurface.
Another example of the use of a treatment train for
creosote-contaminated soil and ground water involves (1)
product removal using a pumping system, (2) flushing with
water and surfactants using pump-and-treat technology, and
(3) in situ biodegradation of the residual contamination (Kuhn
and Piontek, 1989). Each technology is employed in the order
of ease of removal of creosote from the subsurface. The
treatment train selected was based on a site characterization to
identify where the creosote was located and the mass of
creosote (including pure product) associated with subsurface
phases, i.e., the vadose zone and aquifer materials. The prob-
lem assessment identified the following areas of concern: (1)
potential offsite migration of pure product; (2) slow leaching
of low levels of creosote contaminants sorbed to soil, subsur-
face, and aquifer materials; and (3) presence of high molecu-
lar weight polycyclic aromatic compounds that are toxic to
human health, are nonvolatile, and have very low water
solubilities. Each technology was evaluated in laboratory-
scale treatability tests for treatment effectiveness and for case
of application to contaminated materials obtained from the
site. Engineering design and implementation was based on
results of site characterization, mass balance determinations at
the site, and treatability studies,
Information from treatability studies is used to prepare an
approach to the engineering design and implementation of a
remediation system at a specific site that combines the treat-
ment techniques evaluated to construct an appropriate treat-
ment train. The formulation of a treatment train for a site
generally is based upon information from simulations (e.g.,
mathematical modeling) generated from mass balance stud-
ies, treatability studies, and site/soil characterization data.
14.4 Measurement and Interpretation of
Treatment Effectiveness
Typically, subsurface samples are taken from a treatabil-
ity reactor (in situ or prepared bed) from laboratory-, bench-,
or pilot-scale studies, or from a field site. Waste constituents
are extracted from the samples with a solvent or are thermally
desorbed. Compound concentration is usually measured in the
solvent extract or the thermal resorption stream using chemi-
cal instrumentation (e.g., gas or liquid chromatography with
appropriate detectors). This information is termed the "appar-
ent loss" of the compound and refers to the observation that
the compound only has disappeared from the solvent or
extraction phase, but does not necessarily represent a chemi-
cal mass balance (Park et al., 1990). The change in concentra-
tion of the compound in the solvent with time often is used to
calculate rate and extent of decrease in concentration of the
compound in soil. This information is commonly used to
interpret treatment effectiveness for different technologies as
well as to determine engineering strategies and management
approaches, including (1) time required to attain cleanup
target concentrations; and (2) effects of environmental factors
or experimental variables (chemical, physical, or biological)
on treatment effectiveness.
However, additional information is needed to accurately
measure and interpret treatment effectiveness. In order to
understand treatment mechanisms and to base the selection of
treatment technologies on a rational approach, identification
and measurement of distribution among the physical phases
that comprise a subsurface system is necessary. In addition,
the mechanisms by which a compound may be chemically
altered in a subsurface system must be identified and differen-
tiated (Dupont and Reineman, 1986; Goring et al., 1975;
Guenzi, 1974; Park et al., 1988, 1990; Sims et al., 1988;
Stevens et al. 1989; Unterman et al., 1988).
Information obtained about the rate of apparent loss of
chemicals from a subsurface extract can be enhanced with
information about the (1) interphase transfer potential be-
tween solid and gas phases of the subsurface, and (2) knowl-
edge of mechanisms of interactions of compounds with
subsurface phases. This information then provides the basis
for a more rational approach to subsurface remediation. Evalu-
ation of remediation technology effectiveness also can be
based upon specific media (solid, air) and upon specific
mechanisms, such as recovery of the air phase or enhance-
ment of abiotic destruction or biological degradation, to im-
prove treatment. Evaluation of interphase transfer also allows
characterization of routes by which chemicals may migrate
from the subsurface to the multimedia environment that then
may lead to human exposure. Thus, measuring treatment
effectiveness based upon interphase transfer potential (a mass
balance approach) is also valuable for determining risk reduc-
tion and implementing risk management strategies (Park et
al., 1990). The laboratory flask apparatus used for mass
balance determinations (Figure 14-4) also can be used to
measure and compare potential effectiveness for different
treatment scenarios.
211
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14.5 References
Bailey, J.E. and D.F. Ollis. 1986. Biochemical Engineering
Fundamentals, 2nd ed. McGraw-Hill, New York, NY.
Benefield, L. D., J.F. Judkms, and D.L. Weand. 1982. Process
Chemistry for Water and Wastewater Treatment. Prentice-
Hall, Englewood Cliffs, NJ.
Coover, M.P., R.C. Sims, and W.J. Doucette. 1987. Extrac-
tion of Polycyclic Aromatic Hydrocarbons from Spiked
Soil. Journal of the Association of Official Analytical
Chemists 70:1018-1020.
Corbitt, R.A. 1989. Standard Handbook of Environmental
Engineering. McGraw-Hill, New York, NY.
DiGiulio, D.C. and I.H. Suffet. 1988. Effects of Physical,
Chemical, and Biological Variability in Modeling Or-
ganic Contaminant Migration through Soil. In: Superfund
'88, Hazardous Materials Control Research Institute, Sil-
ver Spring, MD, pp. 132-137.
Donigian, A. S., Jr. and P.S.C. Rao. 1986. Overview of Terres-
trial Processes and Modeling. In: Guidelines for Field
Testing Soil Fate and Transport Models, S.C. Hern and
S.M. Melancon, (eds.), EPA/600/4-86/020 (NTIS PB86-
209400), pp. 1-32.
Dupont, R.R. and J.A. Reineman. 1986. Evaluation of Volatil-
ization of Hazardous Constituents at Hazardous Waste
Land Treatment Sites. EPA/600/2-86/071 (NTIS PB86-
233939).
Goring, C.A.I., D.A. Laskowski, J,W. Hamaker, and R.W.
Miekle. 1975. Principles of Pesticide Degradation in Soil.
In: Environmental Dynamics of Pesticides, R. Haque and
W. H. Freek, (eds.), Plenum Press, New York, NY.
Grenney, W.J., C.L. Caupp, R.C. Sims, and T.E. Short. 1987.
A Mathematical Model for the Fate of Hazardous Sub-
stances in Soil: Model Description and Experimental
Results. Hazardous Wastes & Hazardous Materials 4:223-
239.
Guenzi, W.D. (ed). 1974. Pesticides in Soil and Water. Soil
Science Society of America, Madison, WI.
Jury, W.A., W.F. Spencer, and W.J. Farmer. 1983. Behavior
Assessment Model for Trace Organics in Soil: Model
Description. Journal of Environmental Quality 12:558-
564.
Kuhn, R.C. and K.R. Piontek. 1989. A Site-Specific In Situ
Treatment Process Development Program for a Wood
Preserving Site. Paper presented at EPA Technical Pro-
gram on Oily Waste Fate, Transport, Site Characteriza-
tion, and Remediation Seminar, Denver, CO, May 17-18
(Organized by John Matthews, EPA Robert S. Kerr Labo-
ratory, Ada, OK).
Loehr, R. 1989. Treatability Potential for EPA Listed Hazard-
ous Chemicals in Soil. EPA/600/2-89/011 (NTIS PB89-
166581/AS).
Lynch, J. and B.R. Genes. 1989. Land Treatment of Hydro-
carbon Contaminated Soils. In: Petroleum Contaminated
Soils, Vol. 1: Remediation Techniques, Environmental
Fate, and Risk Assessment, P.T. Kostecki and E.J.
Calabrese (eds.), Lewis Publishers, Chelsea, MI, pp. 163-
174,
Mahmood, R.J. 1989. Evaluation of Enhanced Mobility of
PAHs in Soil Systems. Ph.D. Dissertation, Department of
Civil and Environmental Engineering, Utah State Univer-
sity, Logan, UT.
Mahmood, R.J. and R.C. Sims. 1986. Mobility of Organics in
Land Treatment Systems. Journal of Environmental En-
gineering (ASCE) 112:236-245.
McLem, J.E., R.C. Sims, W.J. Doucette, C.L. Caupp, and
W.J. Grenney. 1988. Evaluation of Mobility of Pesticides
in Soil using U.S. EPA Methodology. Journal of Environ-
mental Engineering (ASCE) 114:689-703.
Metcalf and Eddy, Inc. 1979. Wastewater Engineering: Treat-
ment, Disposal, and Reuse. McGraw-Hill, New York,
NY.
Murdoch, L., B. Patterson, G. Losonsky, and W. Harrar. 1988.
Innovative Technologies of Delivery or Recovery: A
Review of Current Research and a Strategy for Maximiz-
ing Future Investigations. EPA/600/2-89/066 (NTIS PB90
156225/AS).
National Academy of Sciences (NAS). 1990. Ground Water
Models: Scientific and Regulatory Applications. National
Academy Press, Washington, DC.
Nkedi-Kizza, P., P.S.C. Rao, and J.W. Johnson. 1983. Ad-
sorption of Diuron and 2,4,5-Ton Soil Particle Separates.
Journal of Environmental Quality 12:195-197,
Park, K. S., R.C. Sims, W.J. Doucette, and J.E. Matthews.
1988. Biological Transformation and Detoxification of
7,12-Dimethylbenz(a) anthracene in Soil Systems. Jour-
nal Water Pollution Control Federation 60:1822-1825.
Park, K. S., R.C. Sims, R.R. Dupont, W.J. Doucette, and J.E.
Matthews. 1990. Fate of PAH Compounds in Two Soil
Types: Influence of Volatilization, Abiotic Loss and Bio-
logical Activity. Environ. Toxicol. Chem. 9:187-195.
Rao, P. S. C., A.G, Hornsby, D.P. Kilcrease, and P. Nkedi-
Kizza. 1985. Sorption and Transport of Hydrophobic
Organic Chemicals in Aqueous and Mixed Solvent Sys-
tems: Model Development and Preliminary Evaluation.
Journal of Environmental Quality 14:376-383.
Rifai, H. S., P.B Bedient, R.C. Bordon, and J.F. Haasbeek.
1989. BIOPLUME II-Computer Model of Two-Dimen-
sional Contaminant Transport Under the Influence of
212
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Oxygen Limited Biodegradation in Ground Water (User's
Manual Version 1.0; Preprocessor Source Code Version
1.0; Source Code Version 1.0). EPA/600/8-88/093 (NTIS
PB89-151 120/AS).
Ross, D., T.P. Marziarz, and A.L. Bourquin. 1988. Bioreme-
diation of Hazardous Waste Sites in the USA: Case
Histories. In: Superfund '88, Hazardous Materials Con-
trol Research Institute, Silver Spring, MD, pp. 395-397.
Scalf, M.R. and D.C. Draper. 1989. RSKERL-Ada Superfund
Technology Support Center The First Two Years. U.S.
Environmental Protection Agency (Robert S. Kerr Envi-
ronmental Research Laboratory, Ada, OK).
Short, T.E. 1986. Modeling Processes in the Unsaturated
Zone. In: Land Treatment A Hazardous Waste Manage-
ment Alternative, R.C. Loehr and J. F. Malina (eds.),
Water Resources Symposium No. 13, University of Texas
Press, Austin, TX, pp. 211-240.
Sims, R.C., D.L. Sorensen, J.L. Sims, J.E. McLean, R.
Mahmood, and R.R. Dupont. 1984. Review of In-Place
Treatment Technologies for Contaminated Surface Soils-
Volume 2: Background Information for In-Situ Treat-
ment. EPA-540/2-84-003b (NTIS PB85-124899).
Sims, R.C., WJ. Doucette, J.E. McLean, W.J. Grenney, and
R.R. DuPont. 1988. Treatment Potential for 56 EPA
Listed Hazardous Chemicals in Soil. EPA/600/6-88/001
(NTIS PB88-174446).
Sims, J.L., R.C. Sims, and J.E. Matthews. 1989. Bioremedia-
tion of Contaminated Soils. EPA/600/9-89/073 (NTIS
PB90-164047).
Sims J.L.. R.C. Sims, and J.E. Matthews. 1990. Approach to
Bioremediation of Contaminated Soils. Hazardous Waste
and Hazardous Materials 7(2): 117-149.
St. John. W.D. andD.J. Sikes. 1988. Complex Industrial
Waste Sites. In: Environmental Biotechnology - Reduc-
ing Risks from Environmental Chemicals through Bio-
technology, G.S. Omenn (ed.), Plenum Press, New
York, NY, pp. 237-252.
Stevens, O.K., W.J. Grenney, and Z. Yan. 1988. User's Manual:
Vadose Zone Interactive Processes Model. Utah State
University, Logan, UT.
Stevens, O.K., W.J. Grenney, Z. Yan, and R.C. Sims. 1989.
Sensitive Parameter Evaluation for a Vadose Zone Fate
and Transport Model. EPA/600/2-89/039 (NTIS PB89-
213987/AS).
Symons, B.D. and R.C. Sims. 1988. Assessing Detoxification
of a Complex Hazardous Waste Using the Microtox™
Bioassay. Archives of Environmental Contamination and
Toxicology 17:497-505.
Symons, B. D., R.C. Sims, and W.J. Grenney. 1988. Fate and
Transport of Organics in Soil: Model Predictions and
Experimental Results. Journal Water Pollution Control
Federation 60:1684-1693.
Unterman, R., D.L. Bedard, M.J. Brennan, L.H. Bopp, F.J.
Mondcllo, R.E. Brooks, D.P. Bobley, J.B. McDerrnotq
C.C. Schwartz, and O.K. Dietnch. 1988. Biological Ap-
proaches for Polychlorinated Biphenyl Degradation. In
Environmental Biotechnology - Reducing Risks from
Environmental Chemicals through Biotechnology, G.S.
Omenn (ed.), Plenum Press, New York, NY, pp. 253-269.
U.S. Environmental Protection Agency (EPA). 1984. Review
of In-Place Treatment Techniques for Contaminated Sur-
face Soils. EPA-540/2-84-003a (NTIS PB85-124881).
U.S. Environmental Protection Agency (EPA). 1986. Permit
Guidance Manual on Hazardous Waste Land Treatment
Demonstrations. EPA-530/SW-86-032 (NTIS PB86-
229184).
U.S. Environmental Protection Agency (EPA). 1988a. Inter-
active Simulation of the Fate of Hazardous Chemicals
during Land Treatment of Oily Wastes: RITZ User's
Guide. EPA/600/8-88-001 (NTIS PB88-195540).
U.S. Environmental Protection Agency (EPA). 1988b. Ground-
water Modeling: An Overview and Status Report. EPA/
600/2-89/028 (NTIS PB89-229497). Also available from
International Ground Water Modeling Center, Butler Uni-
versity, Indianapolis, IN.
U.S. Environmental Protection Agency (EPA). 1989. Guide
for Conducting Treatability Studies under CERCLA. EPA/
540/2-89/058.
U.S. Environmental Protection Agency (EPA). 1990. Hand-
book on In Situ Treatment of Hazardous Waste-Contami-
nated Soils. EPA/540/2-90-W2 (NTIS PB90-155607).
Weaver, J., C.G. Enfield, S. Yates, D. Kreamer, and D. White.
1989. Predicting Subsurface Contaminant Transport and
Transformation: Considerations for Model Selection and
Field Validation. EPA/600/2-89/045 (NTIS PB90-
155615).
Willis, R. and W. W-G. Yeh. 1987. Groundwater Systems
Planning and Management. Prentice Hall, Englewood
Cliffs, NJ.
Wilson, J.T., L.E. Leach, J. Michalowski, S. Vandegrift and
R. Callaway. 1989. In Situ Bioremediation of Spills from
Underground Storage Tanks: New Approaches for Site
Characterization, Project Design, and Evaluation of Per-
formance. EPA/600/2-89/042 (NTIS PB89-219976/AS),
213
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Chapter 15
Remediation Techniques for Contaminated Soils
Ronald C. Sims and Judith L. Sims
Currently, many remedial techniques are being used and
evaluated for cleanup of contaminated soils. Tables 15-1 and
15-2 list participants in the U.S. Environmental Protection
Agency (EPA) SITE program that are testing and evaluating
remedial technologies applicable to contaminated soils (U.S.
EPA, 1989f). Table 15-3 summarizes technologies applicable
to contaminated soils that are currently being demonstrated
and evaluated in the NATO/CCMS Pilot Study, Demonstra-
tion of Remedial Action Technologies for Contaminated Land
andGroundwater (U.S. EPA, 1989d).
Selected physical, chemical, biological, thermal, and fixa-
tion/ encapsulation soil remediation techniques were catego-
rized as in situ and prepared bed and are summarized in Table
15-4 (Rich and Cherry, 1987; U.S. EPA, 1987, 1988c, 1989b).
Each soil remediation technique also was evaluated with
respect to function (separation, detoxification, etc.); potential
for formation of residuals/transformation products; applica-
tions; and limitations. This chapter presents a subset of these
techniques, evaluated at pilot or field scale, that were selected
for additional description.
15.1 In Situ versus Prepared Bed Soil
Remediation
The vadose zone is the region extending from the ground
surface to the upper surface of the principal water-bearing
formation. It is divided into three characteristic areas or belts.
The uppermost belt consists of soil and other materials that lie
near to the surface and discharge perceptible quantities of
water into the atmosphere. The water is discharged by the
action of plants or by soil evaporation and convection. The
lowest belt, the capillary fringe, is located immediately above
the water table and contains water drawn up from the zone of
saturation by capillary action. The intermediate belt lies be-
tween the belt of soil water and the capillary fringe (Lehr,
1988). In this chapter, soil remediation techniques address the
vadose zone and situations where the saturated zone is engi-
neered to become unsaturated, e.g., when ground water is
pumped to create an unsaturated zone.
The two soil treatment processes discussed in this chapter
are in situ treatment and prepared bed treatment. In situ
treatment consists of treating contaminated soil in place, i.e.,
the contaminated soil is not moved from the ground. Mile-
stone publications that should be consulted for scientific and
engineering information specifically addressing in situ treat-
ment include Sims et al. (1984); U.S. EPA (1984); U.S. EPA
(1990); Sims et al. (1989); and Dupont et al. (1988).
In a prepared bed system, the contaminated soil may be
either (1) physically moved from its original site to a newly
prepared area, which has been designed to enhance treatment
and/or to prevent transport of contaminants from the site; or
(2) removed from the site to a storage area while the original
location is prepared for use, then returned to the bed, where
treatment is accomplished. Preparation of the bed may include
placement of a clay or plastic liner to retard transport of
contaminants from the site or addition of uncontaminated soil
to provide additional treatment medium. Treatment may be
enhanced with biological and/or physical/chemical methods,
as with in situ systems (Sims and Sims, 1986; Sims et al.,
1989). Prepared bed treatment approaches are based on modi-
fications of principles developed in the areas of land applica-
tion of solid and liquid wastes and in land treatment of
hazardous wastes (Sims et al., 1989, U.S. EPA, 1983, U.S.
EPA, 1986).
15.2 In Situ Techniques
In situ treatment techniques addressed include (1) soil
vacuum extraction, (2) bioremediation, (3) immobilization,
and (4) mobilization.
15.2.1 Soil Vacuum Extraction (SVE)
Referred to as soil vacuum extraction (SVE), forced air
venting, or in situ air stripping, this technique involves extrac-
tion of air and contaminants from unsaturated soil. In contrast
to a static equilibrium soil system where evaporation of a
chemical is equal to the condensation of the chemical (Figure
15-1), with SVE, clean air is injected or passively flows into
the unsaturated zone. Volatile chemicals then partition from
soil water into soil air, with relative partitioning based on the
air/water partition coefficient (KJ or Henry's Law constant
(Figure 15-2) and the vapor-laden air is removed using vacuum
extraction wells.
Typically, components of SVE consist of vacuum extrac-
tion wells (Figure 15-3), air inlet wells, and vapor monitoring
wells distributed across a contaminated site, and a blower(s)
to control air flow. Extraction wells may be placed vertically
or horizontally, although vertical alignment is typical for
deeper contamination zones and for residues in radial flow
215
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Table 15-1. SITE Demonstration Program Participants with Technologies Applicable to Remediation of Contaminated Soils
Applicable Waste'
Developer
Technology
Inorganic
Organic
American Combustion Technologies, Inc.
Norcross, GA
American Toxic Disposal Inc.
Waukegan, IL
AWD Technologies, Inc.
Burbank, CA
Biotrol, Inc.
Chaska, MN
CF Systems Corporation
Waltham, MA
Chemfix Technologies, Inc.
Metairie, LA
Chemical Waste Management, Inc.
Oakbrook, IL
Dehydro-Tech Corporation
East Hanover, NJ
Ecova Corporation
Redmond, WA
EPOC Water, Inc.
Fresno, CA
Exxon Chemicals, Inc./
RioLinda Chemical Co.
Long Beach, CA
GeoSafe Corporation
Kirkland, WA
HAZCON, Inc.
Brookshire, TX
Horsehead Resources Development Co., Inc.
Monaca, PA
international Waste Technologies/
Gee-Con, Inc.
Wichita, KS
MoTec, Inc.
Austin, Tx
Ogden Environmental Services
San Diego, CA
Ozonics Recycling Corp.
Boca Raton, FL
Resources Conservation Co.
Bellevue, WA
Retech, Inc.
Ukiah, CA
S.M.W. Seiko, Inc.
Redwood City, CA
Shirco Infrared Systems, Inc.
Silicate Technology Corp.
Scottsdale, AZ
Soliditech, Inc.
Houston, TX
Solvent Services, Inc.
San Jose, CA
Terra Vac, Inc.
San Juan, PR
Toxic Treatments (USA) Inc.
San Francisco, CA
Wastach, Inc.
Oak Ridge, TN
Pyreton oxygen burner
Vapor extraction system
integrated vapor extraction
and steam vacuum stripping
Soil washing system
Solvent extraction
Solidification/stabilization
X*TRAX" low temperature
thermal resorption
Carver-Greenfield process
for extraction of oily waste
in situ biological treatment
Leaching and micro filtration
Chemical oxidation/cyanide
destruction
in situ vitrification
Solidification/stabilization
Flame (slagging) reactor
in situ solidification/
stabilization
Liquid/solid contact digestion
Circulating fluidized bed
combustor
Soil washing, catalytic/ozone
oxidation
Solvent extraction (BEST)
Plasma reactor
in situ solidification!
stabilization
infrared thermal destruction
Solidification/stabilization
with silicate compounds
Solidification/stabilization
Steam injection and vacuum
extraction (SIVE)
in situ vacuum extraction
in situ steam/air stripping
Solidification/stabilization
NA
Volatile
NA
Metals
NA
Heavy metals
NA
NA
NA
Specific for
heavy metals
Cyanide
Non-specific
Heavy metals
Heavy metals
Non-specific
NA
NA
Cyanide
NA
Metals
Metals
NA
Metals, cyanide,
ammonia
Metals
NA
NA
NA
Non-specific
radioactive
Non-specific
Volatile and semivolatile
organics includng PCBs,
PAHs, PCPs, some pesticides
Volatile organic compounds
High molecular weight organics
PCBs, volatile, and semivolatile
organic compounds, petroleum
byproducts
High molecular weight organics
Volatile and semivoiatile
organics, PCBs
PCBs, dioxin, oil-soluble
organics
Chlorinated solvents, non-
chlorinatad organic compounds
NA
NA
Non-specific
Not an inhibitor
NA
PCBs, other non-specific
organic compounds
Halogenated and non-
halogenated organic
compounds, pesticides
Halogenatad and non-
halogenated organic
compounds
Semivolatiles, pesticides, PCBs
PCP, dioxin
Specific for high molecular
weight organics
Non-specific
Semivolatile organic
compounds
Non-specific
High molecular weight organics
Non-specific
Volatile and semivolatile
organic compounds
Volatile and semivolatile
organic compounds
Volatile organic compounds
and hydrocarbons
Non-specific
"NA = non applicable
Source: U.S. EPA, 1989f
216
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Table 15-2. SITE Emerging Technology Program Participants with Technologies Applicable to Remediation of Contaminated Soils
Applicable Waste
Developer
Technology
Inorganic
Organic
Babcock & Wilcox Co.
Alliance, OH
Battelle Memorial Institute,
Columbus Division
Columbus, OH
Enviro-Sciences, Inc.
Randolph, NJ
Harmon Environmental Services, Inc.
(formerly Envirite Field Services, Inc.)
Auburn, AL
IT Corporation
Knoxville, TN
Western Research Institute
Laramie, WY
Cyclone combustor
In situ electroacoustic
decontamination
Low energy solvent
extraction
Soil washing
Non-specific
Specific for heavy
metals
NA
NA
Batch steam distillation/ Non-specific
metal extraction
Contained recovery of oily NA
wastes (CROW)
Non-specific
NA
PCBs, other non-specific
organic compounds
Heavy organic compounds
Non-specific
Coal tar derivatives,
petroleum byproducts
NA = non applicable
Source: U.S. EPA, 1989f
patterns (Hutzler, 1990). Schematics of a gas extraction well
and a gas monitoring well are presented in Figures 15-4 and
15-5, respectively.
Important system variables that may affect the perfor-
mance of SVE include properties of the chemical, such as
vapor pressure and volatilization, and properties of the site,
such as soil moisture content, soil texture, and distribution of
contaminants. Vapor pressure is important when a chemical
occurs in a pure phase in the subsurface. Vapor pressures
above 14 mm Hg at 20°C are desirable for application of SVE.
Vapor pressure values for selected subsurface contaminants
are given in Table 15-5. When chemicals are distributed in the
water phase in the soil, the Henry's Law constant is important,
and a dimensionless Henry's constant above 0.01 (mg/L/mg/
L) desirable for use of SVE. Table 15-6 gives Henry's Law
constants for a set of selected organic chemicals where the
application of SVE would be appropriate.
Since movement of volatile organic chemicals (VOCs) is
generally 10,000 times faster in a gas phase than in a water
phase, VOC removal is expected to be enhanced by decreas-
ing soil moisture. However, when soil is very dry, which may
occur when dry air is drawn through soil, VOCs may adsorb
directly onto mineral surfaces, where the magnitude of sorp-
tion is increased and consequently volatilization is decreased
(Figure 15-6). Henry's Law constant is not appropriate under
these conditions, since partitioning is between air and soil
phases only. When moisture is added to soil, the effect is
reversible. The moisture content at which a decrease in vapor
density becomes apparent is often termed the critical moisture
content and generally is equivalent to approximately a mono-
layer of water molecules coating the soil particles (Spencer et
al, 1969, 1973). The effect of soil water content on dieldrin
vapor pressure is illustrated in Figure 15-7. Johnson and
Sterrett (1988) noted that dichloropropane concentrations were
correlated with ambient air moisture during the use of SVE at
a site in Benson, Arizona.
If contaminated soil contains immiscible fluids in the
form of oils, (e.g., petroleum hydrocarbons), the four-com-
partment system discussed previously is operative (water, air,
oil, and soil as discussed in Chapter 14). In this system,
chemical volatility will be affected by the chemical vapor
pressure and mole fraction within the immiscible oil fluid, and
governed by Raoult's Law:
P =X P'
[15-1]
where Pa= vapor pressure of solvent over solution (mm Hg),
Xa= mole fraction of solvent in solution, and P° = vapor
pressure of pure solvent (mm Hg).
For contamination by hydrocarbons with multiple com-
ponents, volatilization will proceed such that lower molecular
weight chemicals will volatilize before higher molecular weight
compounds. Through this process of weathering of the waste/
soil mixture, SVE extraction efficiency is observed to de-
crease to less than 10 percent when the fraction of gasoline
remaining is approximately 40 percent (Figures 15-8 and 15-
9) (Johnson, 1989). Therefore, measuring general parameters
such as total hydrocarbons is not sufficient to indicate the
removal efficiency of individual constituents.
Soil texture has been evaluated as it influences air perme-
ability (DiGiulio et al., 1990). In less permeable media, such
as glacial till and clayey soils, secondary permeability or
porosity (fractures) will dominate air flow. There will be rapid
removal of VOCs in fractures and slow removal in the soil
matrix. In more permeable media, such as sands, sandy loams,
and loamy sands, SVE is appropriate (see Figures 15-10 and
15-11). Pneumatic pump tests in the field are recommended
for site-specific evaluation of SVE application.
Due to release of VOCs from the soil matrix, when
extraction wells are temporarily turned off, concentrations of
VOC increase in soil air (referred to as "VOC rebound ef-
fect"), with an equilibrium concentration that is determined
by Henry's Law constant. When blowers are turned on, an
increase in the concentration of extracted vapor from the soil
will be observed. Diffusive release from subsurface stratigra-
217
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Table 15-3. NATO/CCMS Projects for the Remediation of Contaminated Soils
Treatable Contaminants
Treatment
Organization/site
Aliphatic
Hydrocarbons
Aromatic
Hydrocarbons
Halogenated
Hydrocarbons
Heavy
Metals
Petroleum
Fuels, Oil
Specfic
Contaminants
Treated
Treatment
Location
Status of
Technology
Biological
Enhanced aerobic restoration
U.S. Air Force, Battelle
Eglin Air Force Base, FL
United States
Microbial treatment
Former gas works
Fredensborg, Denmark
Chemical/Physical
K-PEG process
U.S. Environmental
Protection Agency
Wide Beach, NY United States
High pressure soil washing
Scrap metal & copper refinery
Berlin, Federal Republic of Germany
High pressure soil washing and oxidation
Goldbeck Haus, Hamburg
Federal Republic of Germany
Soil vapor extraction
U.S. Environmental Protection Agency
Verona Well Field
Battle Creek, Ml, United States
Stabilization/Solidification
In-situ vitrification
Parsons Chemical Site
Michigan, United States
Electrokinetic
Electro-reclamation
Loppersum
The Netherlands
Thermal
Thermal resorption and destruction
(radiation heating)
Dekonta GmbH, Hamburg
Federal Republic of Germany
Jet fuel
In-situ Experimental
Polycyclic aromatic On-site, Demonstration
hydrocarbons, in-situ
phenols, cyanides
PCBs, dioxin
Lead, PAHCs
Phenol, kresol
On-site, Demonstrated
mobile
On-site Commercial
mobile
In-situ Demonstration
Halogenated and In-situ Demonstrated
aromatic hydrocarbons
Mercury
Arsenic
In-situ Experimental
In-situ Commercial
Chlorobenzenes, On-site Experimental
Chlorophenols,
Hexachlorocyclohexane,
dioxins, furans
Source: U.S. EPA, 1989d
-------
Table 15-4. Selected Remediation Techniques Possibly Suitable for Cleanup of Contaminated Soils
Remediation Techniques
Type of Treatment Treatment Function Possible Residuals/ Possible Applications
Possible Limitations
Technology
Category
Transformation Products
Phyaical/Chemical
Treatment
Low Temperature
Thermal Stripping
(including radio
frequency heating)
Soil Washing
In-tank
In situ
In-tank
Separation
Separation;
volume
reduction
Soil Flushing
In situ
Separation;
volume
reduction
Soil Vacuum Extraction
(SVE)
Glycolate
Dechlorination
In situ
Prepared bed
In-tank
In situ
Separation
Detoxification
Off-gas; spent carbon or
ash from afterburner,
processed soil; hazardous
emissions from in situ
applications
Extracted materials;
water/flushing agent
Extracted materials;
water/washing agent
Volatile organics and
volatile toxic metals
Water/reagent mix;
reaction products
Compounds of low water
volubility and high volatility
Organics and inorganic;
most suitable for soils
contaminated with only a few
specific chemicals
Organics and inorganic;
most suitable for soils
contaminated with only a
few specific chemicals
Volatile organics and toxic
metals; may be enhanced
by the use of steam
Dehalogenation of aromatic
halide compounds
Limited to organics with Henry's Law
constant greater than 3.0 x 10 3atm-m 7
mole and boiling points less than 800°;
more effective for soils with low contents
of organic matter and moisture
Unfavorable contaminant separation
coefficients; less effective with complex
mixtures of waste types and variation
in waste imposition; unfavorable soil
characteristics include: high humic
content, soil/solvent reactions, high
silt and clay content, and clay soils
containing semivolatiles; unfavorable
washing fluid characteristics include:
difficult recovery of solvent or surfactant,
poor treatability of washing fluid,
reduction of soil permeability, and
high toxicity of washing fluid
Unfavorable contaminant separation
coefficients: less effective with complex
mixtures of waste types and variation
in waste composition; unfavorable soil
characteristics include: variable soil
conditions, high organic matter
content, soil/solvent reactions, high
silt and clay content, and clay soils
containing semivolatiles; unfavorable
flushing fluid characteristics include:
difficult recovery of solvent or surfactant,
poor treatability of washing fluid,
reduction of soil permeability, and
high toxicity of washing fluid; requires
containment of leachate and ground water
to prevent off-site groundwater
contamination
Soil heterogeneity (e.g., permeability,
texture); not applicable to saturated
materials or miscible compounds
Heat and excess reagent required for
soils with greater than 20%. moisture
and contaminant concentrations greater
than 5%, and that contain competing
reactive metals (e.g., aluminum)
(continued)
-------
Table 15-4. (Continued)
Remediation Techniques
Type of Treatment Treatment
N>
K>
o
Function
Possible Residuals/
Possible Applications
Possible Limitations
Technology
Neutralization
Oxidation
Photolysis
Precipitation
Reduction
Carbon Adsorption
Ion Exchange
Thermal Treatment
Fluidized Bed
Infrared
pyrolysis
Category
In situ
Prepared bed
In-tank
In situ
Prepared bed
In-tank
Prepared bed
In situ
Prepared bed
In-tank
In situ
Prepared bed
In-tank
In situ
Prepared bed
In situ
Prepared bed
In-tank
In-tank
In-tank
Transformation Products
Detoxification;
immobilization
Detoxification
Detoxification
Separation;
volume reduction;
immobilization
Detoxification
Separation;
immobilization
Separation;
immobilization
Volume
reduction;
detoxification
Volume
reduction:
detoxification
Volume
reduction;
detoxification
Precipitated salts
Oxidized reaction products
Reaction products
Precipitated metals
Reduced reaction products
Processed soil
Processed soil
Off-gases (possibly acidic
and with incomplete com-
bustion products); treated
materials with residual metals;
fly ash: scrubber water
Off-gases (possibly acidic and
with incomplete combustion
products); treated materials with
residual metals; fly ash;
scrubber water
Nonvolatile char and ash (metals,
salts, and particulates)
Waste acids and alkalies
to reduce reactivity and
corrosiveness
Cyanides and
oxidizable organics
Dioxins: nitrated wastes
Metals; certain anions
Chromium, silver, and
mercury
Organic wastes
wastes with high molecular
weight and boiling point and
low volubility and polarity
Metal contaminants
Halogenated and non-
halogenated organics;
inorganic cyanides
Halogenated and non-
halogenated organics;
inorganic cyanides
Wastes not conducive
to conventional incineration;
wastes with volatile metals or
recoverable residues
Compatibility of waste and
treatment chemical to prevent
formation of more toxic or
hazardous compounds
Possible explosive reactions;
production of more toxic or
hazardous products; non-selective
Inability of light to penetrate
soil
Unfavorable effects on soil permeability;
long-term stability unknown
Possible explosive reactions;
production of more toxic or
hazardous products; non-selective
Long-term stability unknown
Selectivity/competition
limitations; pH requirements
High maintenance requirements;
waste size and homogeneity
requirements; applicable
to wastes with low sodium and
metal contents
Limited particle sizes, so may
require size reduction equipment
Small capacity
(conunuea)
-------
Table 15-4. (Continued)
Remediation Techniques
Type of Treatment Treatment
Function
Technology
Rotary Kiln
Category
In-tank
Volume
reduction;
detoxification
Possible Residuals/
Transformation Products
Possible Applications
Possible Limitations
Off-gases (possibly acidic and
with incomplete combustion
products); treated materials with
residual metals; fly ash; scrubber
water
Halogenatad and non-
halogenated organics;
inorganic cyanides
High particulate emissions;
limited particle sizes, so may
require size reduction
equipment
to
Biological Treatment
Aerobic
bioremediation
Anaerobic
bioremediation
Biological Seeding
Comporting
Enzyme addition
In-tank;
prepared bed;
In situ
In-tank;
prepared bed
In situ
In-tank;
prepared bed;
in situ
In-tank;
prepared bed
In-tank;
prepared bed;
In situ
Fixation/Encapsulation
Cement solidification In-tank
In situ
Detoxification Hazardous volatile emissions;
incomplete and possibly
hazardous degradation products;
leachates in soil systems
Detoxification Hazardous volatile emissions;
carbon dioxide, methane and
other gases; incomplete and
possibly hazardous degradation
products; leachates in soil
systems
Detoxification Hazardous volatile emissions;
incomplete and possibly
hazardous degradation products;
leachates in soil systems
Detoxification Hazardous volatile emissions;
incomplete and possibly
hazardous degradation products;
leachates and runoff water
Detoxification Hazardous volatile emissions;
incomplete and possibly
hazardous degradation products;
leachates in soil systems
Storage; Leachates; hazardous volatile
immobilization emissions; solidified waste
materials
Biodegradable organic
wastes
Certain halogenated
organics
Many biodegradable organic
wastes
Biodegradable organic
wastes
Certain biodegradable
organic wastes
Metal cations, latex and
solid plastic wastes
Ability to control environmental
factors conducive to biodegradation;
formation of more toxic or hazardous
transformation products; prepared bed:
area/ limitations due to cost of bed
preparation
May require long treatment
periods; incomplete treatment,
possibly requiring aerobic conditions
to complete degradation process
Survival and activity of organisms in
introduced environment (affected by
environmental factors and competition
with native species)
Maintenance of optimum environmental
conditions for biological activity;
requires large amounts of compost materials
mixed with only about 10% wastes
Activity and stability of introduced enzymes
in natural systems
Incompatible with large amounts of
dissolved sulfate salts or metallic
anions such as arsenates or borates;
setting time increased by presence
of organic matter, lignite, silt, or
clay: requires complete and
uniform mixing of soils and reagents;
long term stability unknown;
may reduce soil permeability and
increase run-off
(continued)
-------
Table 15-4. (Continued)
Remediation Techniques
Type of Treatment Treatment
Technology Category
Function
Possible Residuals/
Transformation Products
Possible Applications
Possible Limitations
Fixation/Encapsulation
Classification/ In-tank
vitification In situ
Storage; Leachates; hazardous volatile
immobilization emissions; glassifie or
vitrified waste materials; aqueous
scrub solution
Inorganic and some
organics in liquids
and contaminated
soils
Long-term stability unknown; high
energy requirements, especially with
high soil water contents and low
permeability; electrical shorting
caused by buried metal drums;
possible underground fire from combustible
materials; volatile metals near
surface may volatilize; site may require
run-off controls
Lime Solidification
(Silicate)
In-tank
In situ
Storage;
immobilization
Leachates, hazardous volatile
emissions; solidified waste
materials
Metals, waste oils, and
solvents
Long-term stability unknown;
incompatible with borates, sulfates,
carbohydrates; requires complete and
uniform mixing of soils and reagents;
may reduce soil permeability and
increase run-off
Thermoplastic In-tank
Microencapsulation In situ
Volume
reduction;
storage;
immobilization
Leachates, hazardous volatile
emissions; encapsulated waste
materials
Complex, difficult to
treat hazardous wastes
Wastes not treatable: wastes with
high water content; strongly
oxidizing contaminants: anhydrous
inorganic salts, tetraborates, iron
and aluminum salts, and organics
with low molecular weights and
high vapor pressures; long-term
stability unknown; requires complete
and uniform mixing of soils and reagents;
may reduce soil permeability and
increase run-off
Sources: Rich and Cherry, 1987; U.S. EPA, 1987, 1988b, 1989b
-------
Unventilated Soil
Soil Water
Figure 15-1. Static soil system in equilibrium (modified from Valsaraj and Thibodeaux, 1988).
Ventilated Soil
Figure 15-2. Enhancement of volatilization through application of soil vacuum extraction (modified from Valsaraj and Thibodeaux,
1988).
223
-------
Vapor
Treatment
Extraction
Well
Air/Water
Separator
Inlet
Wall
| ,
|J
: 1:
: »:
: A:
T
T
t
:Ti
:T:
if:
iT =
:t:
: -t'
:*:
:i-
i*;
: A-:
: Aj
j^j
&ZZZZ2Z2
ft
£ Contaminated &
0 -I IV
, So;/ «
j "
> T-*|4, "
!+-•-
Cap
Tab/e
Figure 15-3. Typical components of a soil vacuum extraction system (from Hutzler et al., 1990).
Threaded Joint
Christy Box @ Grade
Ground Elevation
Slip Cap
4" Diameter PVC Screen
8" Diameter Borehole
Figure 15-4. Schematic of a gas extraction weii used in a soil vacuum extraction system (from DiGiulio, 1989).
224
-------
Ground Elevation
PVC Threaded Cap
2" Diameter
Stainless Steel
1/8" Stainless Steel
4-6" Diameter of Borehole
2" Diameter PVC Vent Pipe
Bentonite Seal
Sand Pack
Figure 15-5. Schematic of a gas monitoring well used in a soil vacuum extraction system (from DiGiulio, 1989).
Table 15-5. Comparative
Constants
Compound
Methylene chloride
Acetone
Methyl ethyl ketone (MEK)
1,2-Dichloroethane (EDC)
Bis (chloromethyl) ether
Phenol
Mercury (Hg°)
PCB-1260
Vapor Pressures
Vapor Pressure
(mm Hg)
362
200
100
61
30.0
0.53
0.0012
4.05 x 10e-5
and Henry's
Henry's Constant
(Dimension/ess)
0.13
miscible
0.001
0.037
0.008
0.00002
0.48
0.30
Table 15-6. Oxygen Supply
Water
Air Saturated
Pure O. Saturated
500mg/IH,Oi
Ib carrier/lb O2
100,000
25,000
10,000
Air
phy of less permeability will cause the slow continual release
of chemicals into the soil-gas phase (Figure 15-12).
Design considerations that affect SVE include extraction
well spacing and extraction well depth. As permeability de-
creases, well spacing decreases; typical well spacings of 10 m
to 30 m are common (Figure 15-13). Also, air circulation
generally is not significant below the screened interval for
extraction wells. Where contamination is deep and permeabil-
ity is high throughout the soil profile, the slotted (screened)
interval should be extended to the maximum depth possible to
maximize treatment, rather than slotted fully vertically (Fig-
ure 15-14).
A promising application of SVE is for enhancement of
biodegradation of volatile and semivolatile chemicals in soils.
SVE provides air to the vadose zone, and thus carries oxygen
that can be used as the terminal electron acceptor (TEA) by
soil microorganisms to biodegrade chemicals (Figure 15-15).
Air has a much greater potential than water for delivering
oxygen to soil on a weight-to-weight and volume-to-volume
basis (Table 15-6). Oxygen provided by air is more easily
delivered since the fluid is less viscous than water higher
oxygen concentrations in air also provide a large driving force
for diffusion of oxygen into less permeable areas within a soil
formation (Miller, 1990).
Hinchee (1989) and Hinchee and Downey (1990) suc-
cessfully applied SVE to enhance biodegradation of petro-
leum hydrocarbons in JP-4 jet fuel at Hill Air Force Base,
Ogden, Utah, by increasing subsurface oxygen concentra-
tions. Soil moisture was found to be a sensitive variable
affecting biodegradation, with increased soil moisture (from
20 percent to 75 percent field capacity) related to increased
biodegradation (Figure 15-16). Monitoring carbon dioxide
and oxygen concentrations, as well as estimating the mass of
VOC biodegraded, is recommended for evaluating potential
enhancement of biodegradation using SVE.
225
-------
Vapor
Phase
Non Polar Organic
HaO
Adsorbed
Layer
Dry
Solid Surface
Wet
Solid Surface
Figure 15-6. Volatile organic carbon adsorption to soil surface in the presence of two soil moisture regimes (from Reible, 1989).
'B
is
f
•
1.0
0.8
0.6
0.4 -
0.2 -
0.0
2.1% Water
3.94% Water
17% Water
20 40 60 SO
Dieldrin in Soil/ppm
100
2.2
2 -
§ 1.8 -
a 1.6 •
1 14-
J«-
1 1'
"§ 0.8 •
I 0.6-
2 0.4 -
0.2 -
0
C7
Total Hydrocarbon
Vapor
20 40 60
Percent Volatilized
80
100
Figure 15-7. Effect of soil water content on dieldrin vapor
pressure (modified from Spencer and Claith,
1989).
Figure 15-8. Volatilization of different hydrocarbon compo-
nents in gasoline (from Johnson, 1989).
226
-------
Total Hydrocarbon
Vapors
Ui
i
0.2
0.4 0.2 0
Fraction Gasoline Remaining
Figure 16-9. Soii vacuum extraction efficiency based on total hydrocarbon vapors (from Johnson, 1989).
Vacuum Extraction Application Related to Soil Texture
Stop and Evaluate Carefully Clay, Silty Clay, Silty Clay Loam
Some Difficulty Sandy Clay, Clay Loam
Good Sandy Clay Loam, Silt Loam
Very Good Sand, Loamy Sand, Sandy Loam,
Loam
Percent by
Weight Clay
Percent by
WeightSilt
Loamy 20
Sand
Sand
100 90 BO 70 60 50 40 30 20 10
Percent by Weight Sand
Figure 15-10. Soil texture trilinear diagram (modified from DiGiulio, 1989).
227
-------
&:>:&:; Sandy Clay ggjggg
Average Velocity
Convection
Diffusion
Convection
Diffusion
Convection
Diffusion &
Convection
Diffusion
Vertical Section through Aquifer
Velocity Profile
Dominant Flow Process
Figure 15-11. Effect of geologic stratification on velocity and resultant dominant flow process (from Keely et ai., in press).
I -J
2
Re-Start Yield Spike
I i
Time —
Figure 15-12. Chemical concentration in the vapor phase versus time for a soil vacuum extraction system where the system is
temporarily discontinued, then restarted (from DiGiulio et al., 1990).
228
-------
f.
3.,
30m
Time (Seconds X 10 6)
Figure 15-13. Effect of well spacing on total solute mass
remaining in soii with vacuum extraction time
(from Wilson et al., 1989).
9m
Time (Seconds X 10°)
Figure 15-14. Effect of weil depth on total solute mass remain-
ing in soil with vacuum extraction time (from
Wilson et al., 1989).
In situ vacuum extraction has been demonstrated in Mas-
sachusetts as part of the Superfund SITE program (U.S. EPA,
1989c, and 1989f), in Michigan and Puerto Rico (U.S. EPA,
1988a), and at several other locations in the United States
(U.S. EPA, 1990).
15.2.2 Bioremediation
Biotic reactions in the subsurface, including definitions
and mechanisms, are addressed in Chapter 13. Wilson (1983)
identified biological processes, including microbial degrada-
tion, as important mechanisms for attenuating contaminants
during transport through the vadose zone to the ground water.
In situ soil remedial measures using biological processes can
reduce or eliminate continuing or potential ground-water con-
tamination, thus reducing the need for extensive ground-water
monitoring and treatment requirements (Wilson, 1981, 1982,
1983).
In situ biological remediation of soils contaminated with
organic chemicals is also an alternative treatment technology
for achieving a permanent cleanup remedy at hazardous waste
Aerobic Biodegradation
Hydrocarbon
Oxygen
+ Nutrients
Biomass
CO., + Hp (Respiration)
Figure 15-15. Aerobic biodegradation using hydrocarbon as the
electron donor and oxygen as the electron
acceptor (from Hinchee, 1989).
sites, as encouraged by the EPA for implementation of the
Superfund Amendments and Reauthorization Act (SARA) of
1986. Information for design of in situ bioremediation is
based on land treatment systems designed for hazardous wastes
(Overcash and Pal, 1979; U.S. EPA, 1983, 1986). These land
treatment designs provide a significant information base for
designing in situ soil remediation systems.
In situ bioremediation involves the use of naturally oc-
curring microorganisms (in contrast to genetically engineered
microorganisms) to degrade and/or detoxify hazardous con-
stituents in the soil at a contaminated site to protect public
health and the environment. Bioremediation techniques for
contaminated soils have been addressed at several scientific
meetings and conferences (AWMA/U.S. EPA, 1989, 1990;
HMCRI, 1989; Omenn, 1988; Lewandowski et al., 1989; U.S.
EPA, 1989a). The use of bioremediation techniques in con-
junction with chemical and physical treatment processes, i.e.,
the use of a "treatment train," is an effective means for
comprehensive site-specific remediation (Ross et al., 1988).
Components of soil bioremediation systems generally
include (1) delivery systems, such as injection nozzles, plows,
and irrigation systems, deliver water, nutrients; oxygen; or-
ganic matter, specialized microorganisms, and/or other amend-
ments, as required; and (2) run-on and run-off controls for
moisture control and waste containment (U.S. EPA, 1984,
1990).
Four approaches are generally used for in situ biological
treatment: (1) enhancement of biochemical mechanisms for
detoxifying or degrading chemicals, (2) augmentation with
exogenous acclimated or specialized microorganisms origi-
nating from uncontaminated or contaminated environments,
(3) application of cell-free enzymes, and (4) vegetative uptake
(U.S. EPA, 1990). Enhancement of biochemical mechanisms
may involve (1) control of soil factors such as contaminant
concentrations that do not severely inhibit microbial activity,
soil moisture, pH, nutrients, and temperature in order to
optimize microbial activity; (2) addition of organic amend-
ments to stimulate cooxidation or cometabolism; (3) control
of soil oxygen by moisture control to accomplish aerobic or
anaerobic biodegradation; and (4) addition of colloidal gas
aphrons (microscopic bubbles of gas) to increase the concen-
tration of terminal electron acceptors (oxygen) in the soil and
229
-------
25% Field Capacity
50% Field Capacity
75% Field Capacity
Sterile Control
Standard Deviation
10
15
20
25
30
35
Figure 15-16. Enhancement of bioremediation of gasoline components using vacuum extraction of soil amended with nutrients and
moisture (from Hinchse, 1989).
thereby enhance aerobic biodegradation (Keck et al., 1989;
Sims et al., 1989; U.S. EPA 1989a, 1990).
The soil contaminant concentration effect on rate and
extent of detoxification of contaminated soil is illustrated in
Figure 15-17. Detoxification of the soil/waste mixture was
measured using the Microtox™bioassay. The Microtox™
assay is an aqueous general toxicity assay that measures the
reduction in light output produced by a suspension of marine
luminescent bacteria in response to an environmental sample
(Bulich, 1979). Bioluminescence of the test organism depends
on a complex chain of biochemical reactions. Chemical inhi-
bition of any of the biochemical reactions causes a reduction
in bacterial luminescence. Therefore, the Microtox™test
considers the physiological effect of a toxicant, not just mor-
tality. Matthews and Bulich (1984) described a method of
using the Microtox™ assay to predict the land treatability of
hazardous organic wastes. Matthews and Hastings (1987)
developed a method using the Microtox™ assay to determine
an appropriate range of waste application loading for soil-
based waste treatment systems. Symons and Sims (1988)
utilized the assay to assess the detoxification of a complex
petroleum waste in a soil environment. The assay also was
included as a recommended bioassay in the EPA's Permit
Guidance Manual on Hazardous Waste Lund Treatment Dem-
onstrations (1986). Comparison of results presented in Figure
15-18 for a clay loam soil with results for the sandy loam soil
shown in Figure 15-17 indicates that detoxification rate and
extent for a waste is a function of soil type. Implications for
management of heavily contaminated soils, therefore, may
include the incorporation of additional treatment medium
(uncontaminated soil) into contaminated soil. This incorpora-
Kldman Sandy Loam
0 2% Oil & Grease
• 4% Oil & Grease
• 8% Oil & Grease
100
BO-
90
Time (days)
120
150 180
Figure 15-17. Detoxification of sandy loam soil measured by
Microtox™assay (from Symonsand Sims, 1988.
230
-------
tion will decrease the concentration of contaminant to levels
that are less inhibitory to soil microbial processes, thereby
rendering treatment more rapidly and completely.
Acclimation of a soil to the presence of a waste is shown
for a fossil fuel-contaminated soil in Table 15-7. The accli-
mated soil was exposed to the fossil fuel waste for one year
before a repeat application of the waste. Results presented in
Table 15-7 indicate that a higher percentage of waste was
treated in the acclimated soil. Treatment also occurred more
rapidly compared to treatment in unacclimated soil. Manage-
ment of contaminated soil, therefore, may include the addition
of lightly contaminated, preexposed, soil to more heavily
contaminated and/or newly contaminated soil to increase the
rate and extent of treatment..
The effect of soil moisture on treatment of contaminated
soil is illustrated in Table 15-8 and Figure 15-19. The chemi-
cal degradation rates given in Table 15-8 indicate more rapid
degradation at a soil moisture content of 60 to 80 percent of
field capacity than at a soil moisture content of 20 to 40
percent. Microtox™ assay results for evaluation of the changes
in toxicity of four wastes (Figure 15-19), two petroleum and
two wood preserving, incubated in relatively dry sandy loam
soil (20 to 40 percent field capacity) over a period of 360 days
indicated little change in toxicity for three wastes and an
increase in toxicity for one waste. Comparison of results
obtained for lower soil moisture (Figure 15-19) with those for
higher soil moisture (Figure 15-17) for petroleum wastes in
sandy loam soil indicate the importance of soil moisture in
influencing microbial activity in waste/soil mixtures.
The effect of temperature on apparent loss of polycyclic
aromatic hydrocarbon (PAHs) compounds in a sandy loam
soil is summarized in Table 15-9 (Coover and Sims, 1987).
Temperature has an important effect on the fate and behavior
of PAHs and, therefore, has implications for seasonal effects
on the rate of biological remediation of soil contaminated with
these chemicals. Microbial ecologists have identified ranges
of critical environmental conditions that affect aerobic activ-
ity of soil microorganisms (Table 15-10). Many of these
conditions are controllable and can be modified to enhance
activity (Huddleston et al, 1986; Paul and Clark, 1989;
Rochkmd et al., 1986; Sims et al., 1984).
The application of cooxidation processes for the biodeg-
radation of high molecular weight PAHs present in oil (NAPL)
phases in soil has been investigated by Keck et al. (1989). In
certain cases, PAH degradation may be limited by the rate of
primary substrate (oil) degradation, which is limited by the
rate of supply of terminal electron acceptors (oxygen) to the
subsurface. In the study by Keck et al., aerobic conditions
were not sufficient to stimulate biodegradation of high mo-
lecular weight PAHs present as a synthetic mixture in soil;
however, when PAHs were present in an oily matrix in the
soil, and the soil was supplied with oxygen, PAHs were
observed to exhibit faster degradation kinetics (Figure 15-20).
Results indicated that oxygen may limit the rate and extent of
biodegradation in soil environments, in addition to saturated
environments. Supplying oxygen to the contaminated vadose
zone may allow biodegradation of oily components of soil
wastes, which may result in simultaneous cooxidation of
resistant PAHs present in the oily waste.
There is also increasing evidence that some halogenated
compounds may be degraded under methanogenic conditions
through a process of reductive dehalogenation (Suflita et al.,
1982, 1983, 1984). Kobayashi and Rittmann (1982) deter-
mined that the redox potential of the environment must be
below 0.35 V for significant reductive dechlorination to oc-
cur. Reductive reactions may be catalyzed by both abiotic and
biochemical means in anaerobic environments.
Oxygen may be consumed faster than it can be replaced
by diffusion from the atmosphere, and the soil may become
anaerobic. Clay content of soil and the presence of organic
matter also may affect oxygen content in soil. Clayey soils
tend to retain a higher moisture content, which restricts oxy-
gen diffusion, while organic matter may increase microbial
activity and deplete available oxygen. Loss of oxygen as a
metabolic electron acceptor induces a change in the activity
and composition of the soil microbial population. Obligate
anaerobic organisms and facultative anaerobic organisms,
which use oxygen when it is present or switch to alternative
electron acceptors such as nitrate or sulfate in the absence of
oxygen, become the dominant populations. Additional infor-
mation concerning in situ anaerobic bioremediation can be
found in the document, Handbook on In Situ Treatment of
hazardous Waste-Contaminated Soils (U.S. EPA, 1990).
The use of plants for stimulating microbial activity in soil
results in increased biodegradation of target organic chemi-
cals in contrast to the possibility of vegetative accumulation
of chemicals for harvesting and removal from a site. This
method is currently being investigated by Walton and Ander-
son (1990) and Aprill and Sims (1990). In soils with low
levels of contamination, plant roots may stimulate the biodeg-
radation of toxic chemicals by providing exudates that serve
as carbon and energy substrates for soil microorganisms. The
effects of prairie grasses on soil PAH concentrations are
summarized i n Table 15-11. For soil with initial concentra-
tions of PAHs of approximately 10 to 50 mg/kg, the presence
of vegetation in the soil (prairie grasses) resulted in a statisti-
cally significant reduction in PAHs, compared with nonveg-
etated soil.
The environmental factors presented in Table 15-10, as
well as waste and soil/site characteristics identified in Chapter
14, interact to affect microbial activity at a specific contami-
nated site. Computer modeling techniques are useful design
and evaluation tools to describe these interactions and their
effects on bioremediation treatment techniques for organic
constituents in a specific situation.
Measurement of physical abiotic loss mechanisms (dis-
cussed in Chapter 13) and partitioning of organic substances
into air and soil phases (discussed in Chapters 10 and 11)
should be used in degradation studies to ensure that generated
information is related to disappearance mechanisms of the
constituents in the soil system (Abbott and Sims, 1989;
Armstrong and Konrad, 1974). This type of information is
needed to more accurately evaluate and select treatment tech-
niques. For example, for organophosphorus pesticides, sorp-
231
-------
Table 15-7 Acclimation of Soil to Complex Foes// Fuel Waste
Unacclimated Soil
Acclimated Soil
PNA
Constituent
Naphthalene
Phenanthrene
Anthracene
Fluoranthene
Pyrene
Benz(a)anthracene
Chrysene
Benz(a)pyrene
Initial Soil
Concentration
(mg/kg-dry wt)
38
30
38
154
177
30
27
10
Reduction in
40 days (%)
90
70
58
51
47
42
25
40
Soil Concentration
after First Reapplication
of Waste (after 168 days
incubation at initial level)
(mglkg-dry wt)
38
30
38
159
160
40
33
12
Reduction in
22 days (%)
100
83
99
82
86
70
61
50
Source: Sims, 1986
Nunn Clay Loam
• 2% Oil & Grease
• 4% Oil & Grease
m 8% Oil & Grease
100
o
&>
r 60-
1
So
Lu
20-
30 60 90 12U 150 180
Time (days)
Figure 15-18. Detoxification of clay loam soil measured by
Microtox™ assay (from Symons and Sims, 1988).
Table 15-8. Effect of Soil Moisture on PNA Degradation
(Reauits Presented as Half-Life in Days)
Moisture Anthracene
20-40% field capacity 43
60-80% field capacity 37
Phenan-
threne
61
54
Fluoran-
thene
559
231
Source: Sims, 1986
tion-catalyzed hydrolysis of ester linkages is known to be an
important influence on soil degradation. An understanding of
abiotic reactions as influenced by sorption and pH of the
system may allow the design of a more effective remediation
strategy. If abiotic controls are not used, the disappearance of
chemicals may be attributed solely to biological activity,
though biological activity may not play the major role in the
degradation of the chemical. Therefore, knowledge of the
reaction mechanism is directly related to efficiency and effec-
tiveness in remediation strategy design and remediation tech-
nique selection.
15.2.3 Immobilization
One way to predict and control the rate of transport of a
constituent through a subsurface system is to describe its
mobility (or relative immobility) by predicting its retardation
(Borden and Bedient, 1987; Mahmood and Sims, 1986). Re-
tardation describes the relative velocity of the constituent
compared to the rate of movement of water through the
subsurface (see Section 10.3 for more information). Retarda-
tion in unsaturated soil can be represented as:
R=l + (rKd/q)
[15-2]
where p = soil bulk density; Kd= soil/water partition coeffi-
cient, which describes the partitioning, between the soil solid
phase and soil water; and 6 = volumetric moisture content. For
a saturated system, 0 is replaced by the porosity of the system.
For additional detail about this process, see Section 10.3.
This information can be used to evaluate treatment tech-
niques for a contaminated soil system (e.g., techniques to
modify the soil/water partition coefficient, such as control of
soil moisture, changes in bulk density, or addition of amend-
ments to the soil). Constituents can be "captured" or contained
within the system by using these techniques, thus allowing
time for degradation at the site or for engineering implementa-
tion and performance of other remediation treatment tech-
niques, such as soil washing (Sims et al, 1989).
Linear retardation of chemicals in the vapor phase is
discussed in Chapter 11. Variables in the equations given in
232
-------
Non-toxic
400
Toxicity of water soluble fraction measured with the Microtox™ assay with incubation time for
PCP-creosote mixed sludge (—13—), creosote sludge (-Q-), API separator sludge (— •-J, and slop c
emulsion solids (-9-) mixed with a Kidman sandy loam soil. EC50 (5, 15°) denotes the effective
concentration (vol/vol) of water soluble extract that reduces light emission of the Microtox™ organism
by 50% five minutes after exposure to the test solution at 15°C. Values presented prior to 250 days o
incubation are the average and standard deviation of duplicate samples. Values presented after 250
days were determined from single samples with a 95% confidence interval.
Figure 15-19. Microtox™ assay results for various materials (from Aprill et al., 1990).
that chapter can be used by professionals involved in treat-
ment technique selection to determine site conditions (pb, Kp,
Ow, 9A) that may influence the effectiveness of specific
treatment technologies. For additional detail about these pro-
cesses, see Section 11.2.2.
Constituents in in situ and prepared bed treatment sys-
tems are generally immobilized through sorption, ion ex-
change, and/or precipitation reactions. These techniques reduce
the rate of contaminant release from the soil environment so
that concentrations along exposure pathways are held within
acceptable limits. The effects of moisture and distribution
coefficient, Kd, on immobilization are illustrated in Figure 15-
21. Results indicate that for chemicals with Kd, values less than
10, management of soil moisture is important with regard to
immobilizing chemicals; for chemicals with Kdvalues greater
than 10, management of soil moisture is less important. Ap-
preaches for controlling soil moisture include run-on and run-
off controls, temporary capping or covering, and irrigation
scheduling.
The cation exchange capacity (CEC) of soil also can be
evaluated with regard to organic as well as metal immobiliza-
tion. Positively charged organic chemicals and metals will
generally readily attach to soil materials with negatively
charged functional groups and negatively charged clay par-
ticles. Addition of clays, synthetic resins, and zeolites will
increase the CEC of soils and increase immobilization of
chemicals sensitive to CEC characteristics of a soil (Sims et
al., 1984; U.S. EPA, 1984). For inorganic chemicals that are
negatively charged in soil systems and can exist in several
oxidation states (e.g., chromium, selenium, and arsenic), im-
mobilization, as well as the toxic form of the chemical, may
potentially be controlled by managing the redox and pH of the
soil system. Management of redox and pH may be short-term
or long-term, depending upon the goals of site management
(e.g., temporary immobilization while delivery and recovery
systems are designed and implemented, followed by soil
flushing with aqueous or surfactant solutions for removal and
recovery of the contaminants) (Sims et al., 1984; U.S. EPA,
1984).
Solidification and stabilization are additional immobili-
zation techniques that are applicable to in situ and prepared
bed systems. These techniques are designed to accomplish
one or more of the following: (1) production of a solid from a
liquid or semisolid waste, (2) reduction of contaminant volu-
bility, and/or (3) a decrease in the exposed surface area across
which transfer may occur. Solidification may involve encap-
sulation of fine waste particles (microencapsulation) or large
blocks of waste (macroencapsulation).
Stabilization refers to the process of reducing the hazard-
ous potential of waste materials by converting contaminants
into their least soluble, mobile, or toxic form (U.S. EPA,
1990). A milestone publication providing additional detail on
this technique is the Handbook for Stabilization Solidifica-
tion of Hazardous Wastes (Cullinane et al., 1986).
Systems for delivering reagents to the contaminated area
include (1) injection systems; (2) soil surface applicators; and/
or (3) delivery and application of electrical energy for melting
-------
Table 15-9. Percentages of PAH Remaining at the End of the 240-Day Study Period and Estimated Apparent Loss Half-Lives
Percent of PAH
Remaining
Compound
Acenaphthene
Fluorene
Phenanthrene
Anthracene
Fluoranthene
Pyrene
Benz[a]anthracene
Chrysene
Benzo[b]fluoran thene
Benzo[k]fluoranthene
Benzo[a]pyrene
Dibenz[a,h]anthracene
Benzofo, h, ijperylene
lndeno[1,2, 3-c, djpyrene
1CFC
5
8
36
83
94
93
82
85
77
93
73
88
81
80
2c
<60
60(+11/-10)
200 (+40/-40)
460(+310/-140)
f
f
680 (+300/-160)
980 (+520/-270)
580 (+520/-180)
910 (+690/-270)
S30(+1700/-230)
820(+1100/-300)
650 (+650/-230)
600 (+31 0/- 150)
2FC
<10
47(+6/-5)
<60
260 (+160/-70)
440 (+560/-1 60)
1900(+6200/-800)
430(+110/-70
1000 (+900/-250)
610 (+590/-200)
1400 (+3300/-560)
290(+570/-120)
750 (+850/-260)
600 (+570/-190)
730(+1100/-270)
3CPC
<10
32 (+S/-3)
<60
200 (+90/-30)
140 (+40/-20)
210 (+160/-60)
240 (+40/-40)
730(+370/-180)
360 (+150/-80)
910 (+4400/-410)
220 (+160/-60)
940(+12000/-450)
590 (+1800/-250)
630 (+2500/-280)
Half lives reported
in the literature (day)
96 "• 45", 0.3-4 '
64", 39", 2-39=
69", 23", 26 c, 9.7", 14"
28", 17", 108-175',
17", 45"
104", 29", 44-182',
39", 34"
73", 27", 3-35 '• 58",
48 «
52", 123", 102-252 c,
240", 130 "
70", 42", 5.5-10.5',
3280,224''
73-130', 85", 65"
143 ", 74 "
91", 69", 30-420 c,
347", 218"
74", 42", 100-190'
179", 70'*
57", 42 ", 200-600'
* t.a (95 percent confidence interval)
" f= 2CPC Sims (1986)
CT = 15-25°C Sims andOvercash (1983)
"T = 20fC PACE (1985)
' T= 2CPC Sims (1982)
' Least squares slope (for calculations of t.J = zero with 95% confidence
Coover and Sims, 1987
Table 15-10 Critical En vlronmental Factors for Microbial Activity
Environmental Factor
Optimum Levels
Available soil water
Oxygen
Redox potential
Nutrients
Temperature
25 - 85% of water holding capacity: -0.01 MPa
Aerobic metabolism: Greater than 0.2 mg/l dissolved oxygen, minimum air-filled
pore space of 10% by volume:
Anaerobic metabolism: Of concentrations less than 1% by volume
Aerobes and facultative anaerobes: greater than 50 millivolts:
Anaerobes: less than 50 millivolts pH 5.5-8.5
Sufficient nitrogen, phosphorus, and other nutrients so not limiting to microbial
growth (Suggested C:N:P ratio of 120:10:1)
15-45°C(Mesophiles)
Sources: Huddleston et al., 1986: Paul and Clark, 1989: Rochkindetal., 1986: Sims et al., 1984
234
-------
Synth, mix
Oil Ref, Waste
Creosote Waste
s
I
vw —
BOO -
600 -
400 -
200 -
0
1 111
if;
3 ' 4
Number of Rings
soils and rocks that contain hazardous materials. Equipment
required for preparing, mixing, and applying reagents de-
pends upon the reagent process, and depth of contamination
(U.S. EPA, 1990).
Important parameters identified by Truett et al. (1983) for
solidification and stabilization of hazardous wastes include
(1) reagent viscosity; (2) permeability of soils; (3) porosity of
waste materials and soil; (4) distribution of waste in surround-
ing material (rocks, soils, etc.); and (5) rate of reaction. The
most significant challenge in applying solidification/stabiliza-
tion treatment in situ is achieving uniform mixing of added
chemical agent(s) with the contaminated soils (U.S. EPA,
1990).
Design factors involve delivery and mixing systems to
obtain complete and uniform distribution of added reagent
throughout the contaminated soil (U.S. EPA, 1990).
In situ solidification/stabilization was applied and evalu-
ated under the Superfund SITE program for treatment of
poly chlorinated biphenyl (PCB) contaminated soils (U.S. EPA,
1989e). Eight additional application sites have been summa-
rized in U.S. EPA (1990).
Figure 15-20. Persistence in soil of PAH compounds as a
function of number of fused benzene rings (from
Keck et si., 1989).
Table 15-11. Apparent Disappearance of PAH Compounds in Vegetated and Unvegetated Soil Using the Tissumizer™ Extraction
Method
Concentration at a Given Time/Initial Concentration (C/Co) (Average ± Std. Dev.)
31
59
94
114*
Benz(a)anthracene
Chrysene
Benzo(a)pyrene
Dibenzfa, h)anthracene
Days of Unvegetated Vegetated Unvegetated Vegetated Unvegetated Vegetated Unvegetated Vegetated
Incubation
0.298*
±0.053
0.150
±0.030
0.089
±0.027
0.063
0.470 •
±0.115
0.144
±0.020
0.061
±0.015
0.064
0.608"
±0.065
0.362
±0.063
0.211"
±0.059
0.124
0.810"
±0.166
0.325
±0.035
0.141"
±0.027
0.133
0.578'
±0.057
0.403
±0.127
0.258
±0.103
0.092
0.895'
±0.124
0.381
±0.086
0.165
±0.033
0.184
1.221 '
±0.131
1.134
±0.240
0.994"
±0.275
0.685
' corresponding values are significantly different with 95% confidence using unpaired t-test comparison.
" corresponding values are significantly different with 90% confidence using unpaired t-test comparison.
c statistics not calculated, only one data point collected.
Aprill and Sims, 1990
1.519'
±0.122
1.011
±0.178
0.663"
±0.099
0.746
151
184
219
0.050'
±0.009
0.067'
±0.024
0.060'
±0.017
0.027'
±0.01 1
0.033'
±0.010
0.027'
±0.014
0.143'
±0.020
0.144"
±0.047
0.119"
±0.045
0.081'
±0.031
0,081"
±0.027
0.064"
±0.019
0.209'
±0.030
0, 188 '
±0.048
0.162"
±0.054
0.122'
±0.045
0.105'
±0.036
0.107"
±0.012
0.949
±0.255
1.011'
±0.090
0.771"
±0.203
0.629
±0.278
0.631'
±0.221
0.547"
±0.048
-------
90
80
70
60.
50
Soil Moisture
• 9 = 20%
* 0 = 40%
o 9 = 60%
• 6 = 80%
W 20
Distribution Coefficient, Kd
30
Figure 15-21. Sorption of chemicals to soil as functions of soil
moisture content and partition coefficient Kd
(Sims et al., 1986).
15.2.4 Contaminant Mobilization
Mobilization of organic ant/or inorganic contaminants
from soil may be accomplished using soil flushing and recov-
ery and treatment of the elutriate (U.S. EPA, 1984, 1990).
Flushing solutions generally include water, acidic and basic
solutions, surfactants, and solvents. The solutions partition a
contaminant into the liquid phase through the volume of
added liquid or by decreasing the distribution coefficient
between the soil and the flushing phase (Sims et al., 1984;
Raghavan et al., 1990). A schematic of a soil flushing system
is shown in Figure 15-22 (U.S EPA, 1984). Components
consist of (1) the flushing solution, and (2) delivery and
recovery systems, which may include injection and recovery
wells, equipment for surface applications, and holding tanks
for storing elutriate for reapphcation (U.S. EPA, 1984, 1990).
Variables affecting application of the technique include
(1) concentration and volume of contamination; (2) distribu-
tion coefficients of waste constituents; (3) interactions of
flushing solutions with soil; and (4) suitability of site for
installation of wells, drains, etc., for delivery and recovery.
Design factors include sizing the delivery and recovery sys-
tems to ensure complete recovery of elutriate. Problems with
respect to flushing of bulk fluids, or NAPLs, from soil sys-
tems are due to the following characteristics of bulk fluids: (1)
low water volubility, (2) high interracial tension, and (3) poor
relative permeability. Relative permeability is defined as:
= [Kd/Ud]/[K,/Uo]
[15-3]
where M = mobility ratio; Kd= fluid permeability (water); K0
= oil permeability; Ud= viscosity of fluid (water); and U0=
viscosity of oil. Strategies for flushing of bulk liquids from
soil generally involve control of one or more of the variables
affecting the mobility ratio through adding chemicals to de-
crease mobility of water or increase mobility of oil (e.g.,
adding surfactants or steam to decrease U0or adding polymers
to increase Ud).
Use of soil flushing in a treatment train with bioremedia-
tion has been evaluated by Dworkin et al. (1988) and by Kuhn
and Piontek (1989) for wood preserving contaminated sites.
Flushing using surfactant/polymer combinations was used to
remove high concentrations of PAH compounds; residual low
concentrations were treated using biological processes.
The effect of adding a solvent on the partitioning of
PAHs between soil and solution (solvent) phases of a soil
system is illustrated in Figure 15-23. When methanol was
used as the solvent in a soil system to flush PAHs from a soil,
the resultant concentration of the PAHs in the solution phase
was several orders of magnitude higher than the concentration
of the PAHs in water.
15.3 Prepared Bed Reactors
In a prepared bed system, the contaminated soil may be
either (1) physically moved from its original site to a newly
prepared area, which has been designed to enhance remedia-
tion and/or to prevent transport of contaminants from the site;
or (2) removed from the site to a storage area while the
original location is prepared for use, then returned to the bed,
where the treatment is accomplished. Preparation of the bed
may include placement of a clay or plastic liner to retard
transport of contaminants from the site or addition of uncon-
taminated soil to provide additional treatment medium.
Possible prepared bed reactor technologies are identified
in Table 15-4 and are evaluated for function as well as
application and limitations. Treatment of contaminants with a
prepared bed may be based on the techniques previously
identified and described for in situ treatment.
An example of the use of a prepared bed reactor for soil
remediation was described by Lynch and Genes (1989). Pre-
pared bed treatment of creosote-contamimted soils from a
shallow, unlined surface impoundment was demonstrated at a
disposal facility for a wood-preserving operation in Minne-
sota. The contaminated soils contained creosote constituents
consisting primarily of PAHs at concentrations ranging from
1,000 to 10,000 ppm. Prior to implementation of the full-scale
treatment operation, bench-scale and pilot-scale studies simu-
lating proposed full-scale conditions were conducted to define
operation and design parameters. Over a 4-month period, 62
to 80 percent removal of total PAHs was achieved in all test
plots and laboratory reactors. Two-ring PAH compounds
were reduced by 80 to 90 percent, 3-ring PAHs by 82 to 93
percent, and 4+-ring PAHs by 21 to 60 percent.
The full-scale system involved preparation of a treatment
area within the confines of the existing impoundment. A lined
waste pile for temporary storage of the sludge and contami-
236
-------
Spray
Application
Pump
D
XT' v^:,-x,^j
X ^....^.'...*.^
Water
Table
Storage
Waste Pond
Well
Leachate
Figure 15-22. Schematic of soil flushing and recycle system (U.S. EPA, 1990).
nated soil from the impoundment was constructed. All stand-
ing water from the impoundment was removed, and the slud-
ges were excavated and segregated for subsequent free oil
recovery. Three to five feet of "visibly" contaminated soil was
excavated and stored in the lined waste pile. The bottom of the
impoundment was stabilized as a base for the treatment area.
The treatment area was constructed by installation of a poly-
ethylene liner, a leachate collection system, 4 feet of clean
backfill, and addition of manure to achieve a carbon: nitrogen
ratio of 50:1. A sump for collection of storm water and
leachate and a center pivot irrigation system also were in-
stalled. The lined treatment area was required because natural
soils at the site were highly permeable. A cap also was needed
for residual contaminants left in place below the liner. Con-
taminated soil was periodically applied to the treatment facil-
ity and rototilled into the treatment soil. Soil moisture was
maintained near field capacity with the irrigation system.
During the first year of operation, greater than 95 percent
reductions in concentration were obtained for 2- and 3-ring
PAHs. Greater than 70 percent of 4- and 5-ring PAH com-
pounds were degraded during the first year. Comparison of
half-lives of PNAs in the full-scale facility were in the low
end of the range of half-lives reported for the test plot units.
Only two PNA compounds were detected in drain tile water
samples, at concentrations near analytical detection limits.
Prepared bed treatment of a Texas oilfield site with
storage pit backfill soils contaminated with styrene, still bot-
tom tars, and chlorinated hydrocarbon solvents was demon-
strated on a pilot scale (St. John and Sikes, 1988). The
remediation efforts included biological, chemical, and physi-
cal treatment strategies. The pilot-scale, solid-phase biologi-
cal treatment facility consisted of a plastic film greenhouse
enclosure, a lined soil treatment bed with an underdrain, an
overhead spray system for distributing water, nutrients, and
inocula, an organic vapor control system consisting of acti-
vated carbon absorbers, and a fermentation vessel for prepar-
ing microbial inoculum or treating contaminated leachate
from the backfill soils. Soils were excavated from the con-
taminated area and transferred to the treatment facility. Aver-
age concentrations of volatite organic compounds (VOCs)
were reduced by more than 99 percent during the 94-day
period of operation of the facility; most of the removal was
attributed to air stripping. Biodegradation of semivolatile
compounds reduced average concentrations by 89 percent
during the treatment period.
237
-------
1
10 -
x^_x Fluoranthene S • 0.38 C (r2- 0.997)
£*--^ Anthracene S • 0.27 C (r2• 0.990
D—D Benzo (o) pyrene S • 1.31 C °-60(r2- 0.997)
O O Indeno (1,2,3 • cd) pyrene S • 1.5 C °-74 (r ?• 0.984)
10
20 30 40
Concentration (mg/L)
50
60
Figure 15-23. PAH adsorption isotherms with methanol and clay loam soi\ (from Mahmood and Sims, 1985).
15.4 References
Abbott, C. and R.C. Sims. 1989. Use of Bioassays to Monitor
Polycyclic Aromatic Hydrocarbon Contamination in Soil.
In: Superfund '89, Hazardous Materials Control Research
Institute, Silver Spring, MD, pp. 23-26.
Aprill, W. and R.C. Sims. 1990. Evaluation of the Use of
Prairie Grasses for Stimulating Polycyclic Aromatic Hy-
drocarbon Treatment in Soil. Chemosphere 20:253-265.
Aprill, W., R.C, Sims, J.L. Sims, and J.E. Matthews. 1990.
Assessing Detoxification and Degradation of Wood Pre-
serving and Petroleum Wastes in Contaminated Soil.
Waste Manag. and Res. 8:45-65.
Armstrong, D.E. and J.G. Konrad. 1974. Nonbiological Deg-
radation of Pesticides. In: Pesticides in Soil and Water,
W. D. Guenzi (ed.), Soil Science Society of America,
Madison, WI, Chapter 7.
AWMA/EPA. 1989. Proceedings of the International Sympo-
sium on Hazardous Waste Treatment: Biosystems for
Pollution Control (Cincinnati, OH). Air and Waste Man-
agement Association, Pittsburgh.
AWMA/EPA. 1990. Proceedings of the International Sympo-
sium on Hazardous Waste Treatment: Treatment of
Contaminated Soils (Cincinnati, OH). Air and Waste
Management Association, Pittsburgh.
Borden, R.C., P.B. Bedient, M.D. Lee, C.H. Ward, and J.T.
Wilson. 1986. Transport of Dissolved Hydrocarbons In-
fluenced by Reaeration and Oxygen Limited Biodegrada-
tion. II. Field Application. Water Resources Research
22:1983-1990.
Bulich, A.A. 1979. Use of Luminescent Bacteria for Deter-
mining Toxicity in Aquatic Environments. In: Aquatic
Toxicology, L.L. Markings and R.A. Kimerle, (eds.),
American Society for Testing and Materials, Philadel-
phia, PA, pp. 98-106.
Coovcr, M.P. and R.C. Sims. 1987. The Effect of Tempera-
ture on Polycyclic Aromatic Hydrocarbon Persistence in
an Unacclimated Soil. Hazardous Waste and Hazardous
Matenals 4:69-82.
Cullinane, Jr., M.J., L.W. Jones, and P.O. Malone. 1986.
Handbook for Stabilization/ Solidification of Hazardous
Wastes. EPA/540/2-86/001 (NTIS PB87-116745/REB).
DiGiulio, D.C. 1989. U.S. EPA Robert S. Kerr Environmental
Research Laboratory, Ada, OK, personal communica-
tion.
DiGiulio, D.C., J.S. Cho, R.R. DuPont, and M.W. Kemblowski.
1990. Conducting Field Tests for Evaluation of Soil
Vacuum Extraction Application. In: Proc. 4th Nat. Out-
door Action Conf. on Aquifer Restoration, Ground Water
Monitoring and Geophysical Methods, National Water
Well Association, Dublin, OH, pp. 587-601.
Dupont, R. R., R.C. Sims, J.L. Sims, andD.L. Sorensen. 1988.
In Situ Biological Treatment of Hazardous Waste-Con-
taminated Soils. In: Biotreatment Systems, D. L. Wise
(ed.), CRC Press, Boca Raton, FL, Chapter 2.
Dworkin, D., D.J. Messinger, and R.M. Shapot. 1988. In Situ
Flushing and Bioreclamation Technologies at a Creosote-
Based Wood Treatment Plant. In: Proc. 5th National
Conference on Hazardous Wastes and Hazardous Materi-
als, Hazardous Materials Control Research Institute, Sil-
ver Spring, MD, pp. 67-78.
Hinchee, R. 1989. Enhanced Biodegradation Through Soil
Venting. Presented at Workshop on Soil Vacuum Extrac-
238
-------
tion held at US. EPA Robert S. Kerr Environmental
Research Laboratory, Ada, OK, April 27-28 (Dominic
DiGiulio, Technical Coordinator).
Hinchee R., and D. Downey. 1990. In Situ Enhanced Biodeg-
radation of Petroleum Distillates in the Vadose Zone. In:
Proceedings of the International Symposium on Hazard-
ous Waste Treatment: Treatment of Contaminated Soils
(Cincinnati, OH). Air and Waste Management Associa-
tion, Pittsburgh, PA.
Hazardous Materials Control Research Institute (HMCRI).
1989. Symposium on Use of Genetically Altered or
Adapted Organisms in the Treatment of Hazardous Wastes.
HMCRI, Silver Spnng, MD.
Huddleston, R.L., C.A. Bleckman, and J.R. Wolfe. 1986.
Land Treatment Biological Degradation Processes. In:
Land Treatment A Hazardous Waste Management Alter-
native, R.C. Loehr and J.F. Malina (eds.), Water Re-
sources Symposium No. 13, University of Texas Press,
Austin, TX, pp. 41-61.
Hutzler, N.J., B.E. Murphy, and J.S. Gierkc. 1990. State of
Technology Review: Soil Vapor Extraction Systems. EPA/
600/2-89/024 (NTIS PB89-195 184).
Johnson, R.L. 1989. Soil Vacuum Extraction: Laboratory and
Physical Model Studies. Presented at Workshop on Soil
Vacuum Extraction held at U.S. EPA Robert S. Kerr
Environmental Research Laboratory, Ada, OK, April 27-
28 (Dominic DiGiulio, Technical Coordinator),
Johnson, J.J. and R.J. Sterrett. 1988. Analysis of In Situ Air
Stripping Data. In: Proc. 5th National Conference on
Hazardous Waste and Hazardous Materials, Hazardous
Materials Control Research Institute, Silver Spring MD,
pp. 451-455.
Keck, J., R.C. Sims, M. Coorer, K. Park, and B. Symons.
1989. Evidence of CoOxidation of Polynuclear Aromatic
Hydrocarbons in Soil. Water Research 23(12):1467-9476.
Keely, J.W., D.C. Bouchard, M.R Scalf, and C.G. Enfield.
Practical Limits to Pump and Treat Technology for Aqui-
fer Remediation. Ground Water Monitoring Review. In
press.
Kobayashi, H. and B.E, Rittmann. 1982. Microbial Removal
of Hazardous Organic Compounds. Environ. Sci. Tcchnol.
16:170A-183A.
Kuhn, RC. and K.R Piontek. 1989. A Site-Specific In Situ
Treatment Process Development Program for a Wood
Preserving Site. Presented at Seminar on Oily Waste
Fate, Transport, Site Characterization, and Remediation,
Denver, CO, May 17-18 (John Matthews, Technical Co-
ordinator, Robert S. Kerr Environmental Research Labo-
ratory, Ada, OK).
Lehr, J. 1988. The Misunderstood World of Unsaturated
Flow. Ground Water Monitoring Review 8(2):4-6.
Lewandowski, G., P. Armenante, and B. Baltzis (eds.). 1989.
Biotechnology Applications in Hazardous Waste Treat-
ment. Engineering Foundation, New York, NY.
Lynch, J, and B.R. Genes. 1989, Land Treatment of Hydro-
carbon Contaminated Soils. In: Petroleum Contaminated
Soils, Vol. I: Remediation Techniques, Environmental
Fate, and Risk Assessment, P.T. Kostecki and E.J.
Calabrese (eds.), Lewis Publishers, Chelsea, MI, pp. 163-
174.
Mahmood, R.J. and RC. Sims. 1985. Enhanced Motility of
Polynuclear Aromatic Compounds in Soil Systems. In:
Proc. 1985 Environmental Engineering Specialty Annual
Conference (Boston, MA), American Society of Civil
Engineers, pp. 128-135.
Mahmood, R.J. and R.C. Sims. 1986. Mobility of Organics in
Land Treatment Systems. Journal of Environmental En-
gineering (ASCE) 112:236-245.
Matthews, J.E. and A.A. Bulich. 1984. A Toxicity Reduction
Test System to Assist Predicting Land Treatability of
Hazardous Wastes. In: Hazardous and Industrial Solid
Waste Testing: Fourth Symposium, J.K. Petros, Jr., W.J.
Lacy, and R.A. Conway (eds.), ASTM STP-886, Ameri-
can Society of Testing and Materials, Philadelphia, PA,
pp. 176-191.
Matthews, J.E. and L. Hastings. 1987. Evaluation of Toxicity
Test Procedure for Screening Treatability Potential of
Waste i. Soil. Toxicity Assessment: International Quar-
terly 2:265-21.
Miller, R. 1990. A Field Scale Investigation of Enhanced
Petroleum Hydrocarbon Biodegradation in the Vadose
Zone Combining Soil Venting as an Oxygen Source with
Moisture and Nutrient Addition. Ph.D. Dissertation, De-
partment of Civil and Environmental Engineering, Utah
State University, Logan, UT.
Omenn, G.S. (ed.). 1988. Environmental Biotechnology—
Reducing Risks from Environmental Chemicals through
Biotechnology. Plenum Press, New York, NY, 505 pp.
Ovcrcash, M.R. and D. Pal. 1979. Design of Land Treatment
Systems for Industrial Wastes—Theory and Practice. Ann
Arbor Science, Ann Arbor, MI.
PACE. 1985. The Persistence of Polynuclear Aromatic Hy-
drocarbons in Soil. PACE Report No. 85-2. Petroleum
Association for Conservation of the Canadian Environ-
ment, Ottawa, Ontario, Canada.
Paul, E.A. and F.E. Clark. 1989. Soil Microbiology and
Biochemistry. Academic Press, San Diego, CA.
Raghavan, R., E. Coles, and D. Dietz. 1990. Cleaning Exca-
vated Soil Using Extraction Agents: A State-of-the-Art
Review. EPA/600/2-89/034 (NTIS PB89-212757/AS).
239
-------
Reible, D.D. 1989. Introduction to Physicochemical Processes
Influencing Enhanced Volatilization. Presented at Work-
shop on Soil Vacuum Extraction held at U.S. EPA Robert
S. Kerr Environmental Research Laboratory, Ada, OK,
April 27-28 (Dominic DiGiulio, Technical Coordinator).
Rich, G. and K. Cherry. 1987. Hazardous Waste Treatment
Technologies. Pudvan Publishing, Northbrook, IL, Chap-
ter?.
Rochkmd, M.L., J.W. Blackburn, and G. Sayler. 1986. Micro-
bial Decomposition of Chlorinated Aromatic Compounds.
EPA/600/2-86/090 (NTIS PB87-116943/REB).
Ross, D., T.P. Marziarz, and A.L. Bourquin. 1988. Bioreme-
diation of Hazardous Waste Sites in the USA: Case
Histories. In: Superfund '88, Hazardous Materials Con-
trol Research Institute, Silver Spring, MD, pp. 395-397.
Sims, R.C. 1982. Land Treatment of Polynuclear Aromatic
Compounds. Ph.D Dissertation, Department of Biologi-
cal and Agricultural Engineering, North Carolina State
University, Raleigh, NC.
Sims, R.C. 1986. Loading Rates and Frequencies for Land
Treatment Systems. In: Land Treatment—A Hazardous
Waste Management Alternative, R.C. Loehr, and J.F.
Malina (eds.), Water Resources Symposium No. 13, Uni-
versity of Texas Press, Austin, TX, pp. 151-170.
Sims, R.C., and M.R. Overcash. 1983. Fate of Polynuclear
Aromatic Compounds (PNAs) in Soil-Plant Systems. Resi-
due Reviews 86:1-68.
Sims, R.C. and J.L. Sims. 1986. Cleanup of Contaminated
Soil. In: Utilization, Treatment, and Disposal of Waste on
Land, K.W. Brown, B.L. Carlile, R.H. Miller, E.M.
Turledge, and B.C.A. Runge (eds.), Soil Science Society
of America, Madison, WI, pp. 257-277.
Sims, R.C., D.L. Sorensen, J.L. Sims, J.E. McLean, R.
Mahmood, and R.R. DuPont. 1984. Review of In-Place
Treatment Technologies for Contaminated Surface Soils-
Volume 2: Background Information for In-Situ Treat-
ment. EPA/540/2-84-003b (NTIS PB85-124899).
Sims, R.C., D. Sorensen, J.L. Sims, J. McLean, R.J Mahmood,
R. Dupont, and J. Jurinak. 1986. Contaminated Surface
Soils In-Place Treatment Techniques. Pollution Technol-
ogy Review No. 132. Noyes Publications, Park Ridge,
NJ, 536 pp.
Sims, J. L., R.C. Sims, and J.E. Matthews. 1989. Bioremedia-
tion of Contaminated Soils. EPA/600/9-89/073 (NTIS
PB90-164047).
Spencer, W.F. and MM. Cliath. 1969. Vapor Density of
Dieldrin. Environ. Sci. Technol. 3:670-674.
Spencer, W.F., M.M. Cliath, and W.J. Farmer. 1969. Vapor
Density of Soil-Applied Dieldrin as Related to Soil-
Water Content, Temperature and Dieldrin Concentration.
Soil Sci. Soc. Am. Proc. 33:509-511.
Spencer, W .F., W.J. Farmer, and M.M. Cliath. 1973. Pesticide
Volatilization. Residue Reviews 49:1-47.
St. John, W. D., and D.J. Sikes. 1988. Complex Industrial
Waste Sites. In: Environmental Biotechnology-Reduc-
ing Risks from Environmental Chemicals through Bio-
technology, G.S. Omenn (ed.), Plenum Press, New York,
NY, pp. 237-252.
Suflita, J. M., A. Horowitz, D.R. Shelton, and J.M. Tiedje.
1982. Dehalogenatiom a Novel Pathway for the Anaero-
bic Biodegradation of Haloaromatic Compounds. Sci-
ence 218:1115-1117.
Suflita, J.M., J.A. Robinson, and J.M. Tiedje. 1983. Kinetics
of Microbial Dchalogenation of Haloaromatic Substrates
in Methanogenic Environments. Appl. Environ. Microbiol.
45:1466-1473.
Suflita, J. M., J. Stout, and J.M. Tiedje. 1984. Dechlorination
of (2,4,5-Trichlorophenoxy) Acetic Acid by Anaerobic
Microorganisms. Journal of Agricultural and Food Chem-
istry 32:218-221.
Symons, B.D. and R.C. Sims. 1988. Detoxification of a Com-
plex Hazardous Waste Using the Microtox™Bioassay.
Archives of Environmental Contamination and Toxicol-
ogy 17:497-505.
Truett, J. B., R.L. Holbergcr, and K.W. Barrett, 1983. Feasibil-
ity of In Situ Solidification/Stabilization of Landfilled
Hazardous Wastes. EPA/600/2-83/088 (NTIS PB83-
261099).
U.S. Environmental Protection Agency (EPA). 1983. Hazard-
ous Waste Land Treatment & EPA SW-874.
U.S. Environmental Protection Agency (EPA). 1984. Review
of In-Place Treatment Techniques for Contaminated Sur-
face Soils. EPA/540/2-84-003 a (NTIS PB85-124881).
U.S. Environmental Protection Agency (EPA). 1986. Permit
Guidance Manual on Hazardous Waste Land Treatment
Demonstrations. EPA/530/SW-86-032 (NTIS PB86-
229 184).
U.S. Environmental Protection Agency (EPA). 1987. A Com-
pendium of Technologies Used in the Treatment of Haz-
ardous Wastes. EPA/625/8-87/014 (Available from Center
for Environmental Research Information, Cincinnati, OH).
U.S. Environmental Protection Agency (EPA). 1988a. Cleanup
of Releases from Petroleum USTS: Selected Technolo-
gies. EPA/530/UST-88/001 (NTIS PB88-241856).
U.S Environmental Protection Agency (EPA). 1988b. Tech-
nology Screening Guide for Treatment of CERCLA Soils
and Sludges. EPA/540/2-88/004 (NTIS PB89-132674/
REB).
240
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U.S. Environmental Protection Agency (EPA). 1989a. Biore-
mediation of Hazardous Waste Sites Workshop: Speaker
Slide Copy and Supporting Information. EPA CERI-89-
11 (NTIS PB89-169205/REB).
U.S. Environmental Protection Agency (EPA). 1989b. Cor-
rective Action: Technologies and Applications. EPA/
625/4-89/020 (Available from Center for Environmental
Research Information, Cincinnati, OH).
U.S. Environmental protection Agency (EPA). 1989c. Dem-
onstration Bulletin: In Situ Vacuum Extmct.ion, Terra
Vac, Inc. EPA/540/M5-89/O03 (NTIS PB90-126665/
GAR).
U.S Environmental Protection Agency (EPA). 1989d. Dem-
onstration of Remedial Action Technologies for Con-
taminated Land and Groundwater. In: Proceedings and
Appendices, Third International Conference, NATO Com-
mittee on Challenges of Modern Society (CCMS)
(Montreal, Canada), pp. v-vii.
U.S. Environmental Protection Agency (EPA). 1989e. Tech-
nology Evaluation Report: SITE Program Demonstration
Test, International Waste Technologies, In Situ Stabiliza-
tion/Solidification, Hialeah, Florida, Volume 1. EPA/
540/5-89/004a (NTIS PB89-194161/AS).
U.S. Environmental Protection Agency (EPA). 1989f. The
Superfund Innovative Technology Evaluation Program:
Technology Profiles. EPA/540/5-89/013 (NTIS PB90-
249756/A07).
U.S. Environmental Prelection Agency (EPA). 1990. Hand-
book on In Situ Treatment of Hazardous Waste-Contami-
nated Soils. EPA/540/2-90-002 (NTIS PB90-155607).
Valsaraj, K.T. and L.J. Thibodcaux. 1988. Equilibrium Ad-
sorption of Chemical Vapors on Surface Soils, Landfills,
and Landfarms-A Review. J. Hazardous Materials 19:79-
99.
Wahon, B.T., and T.A. Anderson. 1990. Microbial Degrada-
tion of Trichlorocthylene in the Rhizosphere: Potential
Application to Biological Remediation of Waste Sites.
Appl. Environ. Microbiol. 56:1012-1016.
Wilson, L.G. 1981. Monitoring in the Vadose Zone: Part I,
Storage Changes. Ground Water Monitoring Review
1(3):32-41.
Wilson, L.G. 1982. Monitoring in the Vadose Zone: Part II.
Ground Water Monitoring Review 2(2):31-42.
Wilson, L.G. 1983. Monitoring in the Vadose Zone: Part III.
Ground Water Monitoring Review 3(1): 155-166.
Wilson, D.J., R.O., Mutch, Jr., and A.N. Clarke. 1989. Model-
ing of Soil Vapor Stripping. Presented at Workshop on
Soil Vacuum Extraction held at U.S. EPA Robert S. Ken-
Environmental Research Laboratory, Ada, OK, April 27-
28 (Dominic DiGiulio, Technical Coordinator).
241
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Chapter 16
Aquifer Restoration
Ronald C. Sims and Judith L. Sims
Currently, several remedial techniques are being used to
restore contaminated ground water and aquifer material. Par-
ticipants in the U.S. Environmental Protection Agency's (EPA)
SITE program that are testing technologies applicable to
contaminated ground water are listed in Tables 16-1 and 16-2
(U.S. EPA, 1989a). Table 16-3 summarizes technologies ap-
plicable to contaminated ground water currently being evalu-
ated and demonstrated in the NATO/CCMS Pilot Study on
demonstration of remedial action technologies for contami-
nated land and ground water (U.S. EPA, 1989b).
The pattern of contamination from a release of contamin-
ants into the subsurface environment, such as would occur
from an underground leaking storage tank containing non-
aqueous phase liquids (NAPLs), is complex (Figure 16-1)
(Palmer and Johnson, 1989; Wilson et al, 1989). As contami-
nants move through the unsaturated zone, a portion is left
behind, trapped by capillary forces. If the release contains
volatile contaminants, a plume of vapors forms in the soil
atmosphere in the vadose zone. If the release contains NAPLs
less dense than water (LNAPLs), they may flow by gravity
down to the water table and spread laterally. Ground water
moving through subsurface sediments contacts the release and
the more water-soluble components are dissolved into the
water phase. Therefore, three distinct regions of contaminants
are formed in the release: a plume of fumes in the soil
atmosphere, a ground-water plume, and the region that con-
tains the oily phase material that serves as the source area for
both plumes. This latter region may include both recoverable
free product (i.e., continuous phase material), and sorbed or
capillary-held material (i.e., residual saturation material). If
the release contains DNAPLs, these contaminants can pen-
etrate to the bottom of an aquifer, forming pools in depres-
sions.
This chapter discusses three techniques concerning aqui-
fer restoration: (1) product removal, (2) pump-and-treat, and
(3) biorestoration.
16.1 Product Removal
Product removal generally consists of product character-
ization, product location, and product recovery. Product char-
acterization refers to identifying the type of product (e.g.,
petroleum, wood-preserving, or solvent) and associated indi-
vidual chemicals (e.g., BXT, PAHs, TCE). Product locations
include characterizing the product mobility at the site (e.g.,
LNAPL following the water table, DNAPL following the
bedrock). Knowledge of whether the product is an LNAPL or
a DNAPL may help locate the product in the subsurface.
Physical recovery techniques to remove free product
include (1) a single pump system producing a mixture of
hydrocarbon and water that must be separated, but requiring
minimal equipment and drilling; (2) a two-pump, two-well
system utilizing one pump to produce a water table gradient
and a second well to recover floating product; or (3) a single
well with two pumps in which a lower pump produces a
gradient and an upper pump collects free product (Lee and
Ward 1986). Vacuum extraction of volatilizing contaminants
also may be used to recover floating free product from a
perched water table.
Pumping systems commonly used for recovery of LNAPLs
arc shown in Figures 16-2 and 16-3. An aboveground oil/
water separator generally is used to recover product for future
use. Subsurface drains also have been used for recovery of
DNAPLs (Figure 16-4a). When only the oil recovery drainline
(ORD) is used (Figure 16-4b), water truncates the flow of
product (DNAPL) due to the poor relative permeability of the
product as described previously in the discussion of soil
flushing. The water table depression drainline (WTDD) is an
efficient method (see Figure 16-4c) to drag an oily product
across the subsurface by viscous forces and thereby create a
hydraulic head of oil above the ORD; however, oil also enters
the WTDD, thereby creating the need for aboveground sepa-
ration of product and water. When both ORD and WTDD are
used (Figure 16-4d), subsurface separation of oil and water is
achieved, thereby minimizing aboveground separation re-
quirements. This system (Figure 16-4d) is also efficient since
the permeability of oil is greatest in the oily contaminated
subsurface, and the underground separation maintains water
flowing in the water compartment and oil flowing in the oily
compartment.
Caution should be exercised during product recovery of
LNAPL when an extraction well is used to control local
gradients and collect free product in a cone of depression. Due
to capillary forces in the subsurface aquifer material, trapped
residual will constitute a continuous source of contamination
to ground water that will persist after product removal from
the water table is completed.
243
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Table 16-1.
SITE Demonstration Program Participants with Technologies Applicable to Remediation of Contaminated Ground
Water
Applicable Waste
Developer
Technology
Inorganic
Organic
AWD Technologies, Inc.
Burbank, CA
(004)
Biotrol, Inc.
Chaska, NM
(003)
DETOX, inc.
Dayton, OH
(003)
E.I. Du Pont de Nemours
and Co./Oberiln Filter Co.
Newark, DE
(003)
Ecova Corporation
Redmond, WA
(003)
Exxon Chemicals, Inc./
Rio Linda Chemical Co.
Long Beach, CA
(004)
Freeze Technologies Corp.
Raleigh, NC
(003)
Ozonics Recycling Corp.
Boca Raton, FL
(004)
Silicate Technology Corp.
Scottsdale, AZ
(003)
Ultrox International, Inc.
Santa Ana, CA
(003)
Zimpro/Passavant,
Inc., Rothschild, Wl
(002)
Integrated Vapor Extraction
and Steam Vacuum
Stripping
Biological Aqueous
Treatment Systems
Submerged Aerobic Fixed-
Film Reactor
Membrane Microfiltration
In Situ Biological Treatment
Chemical
Oxidation/Organics
Destruction
Freezing Separation
Soil Washing, Catalytic/
Ozone Oxidation
Solidification/Stabilization
with Silicate Compounds
Ultraviolet Radiation and
Ozone Treatmen!
PACr/Wet Air Oxidation
NA
Can be applied to Nitrates
Metals inhibit process
Heavy Metals, Cyanide,
Uranium
NA
NA
Non-specific
Cyanide
Metals, Cyanide, Ammonia
NA
NA
Volatile Organic Compounds
Chlorinated and
Nonchlorinated
Hydrocarbons
Readily Biodegradable
Organic Compounds
Non-specific
Chlorinated Solvents,
Nonchlorinated Organic
Compounds
Non-specific
Non-specific
Semivolatiles, Pesticides,
PCBs, POP, Dioxin
High Molecular Weight
Organics
Halogenated Hydrocarbons,
Volatile Organic Compounds,
Pesticides, PCBs
Volatile and Semivolatile
Organic Compounds
NA = non applicable
U.S. EPA, 1989a
16.2 Pump-and-Treat Remediation
Both hydrogeologic information and contaminant infor-
mation are required for pump-and-treat remediation. Hydro-
geologic information about ground-water flow includes
geological and hydraulic factors (described in Chapters 3 and
4) as well as ground-water use/withdrawal factors.
When pump-and-treat remediation is selected, a decision
needs to be made about the use of wells or drains (U.S. EPA,
1990). If the hydraulic conductivity is sufficiently high to
allow flow to wells, then wells are recommended. For low-
permeability material, drains may be required. Wells can be
categorized as extraction, injection, or a combination. Injec-
tion wells reduce cleanup time required by flushing chemicals
to the extraction wells. Design and management decisions
concerning extraction wells include whether to use continu-
ous pumping, pulsed pumping, or pumping combined with
containment. While continuous pumping maintains an inward
hydraulic gradient, pulsed pumping allows maximum concen-
trations to be pumped and requires only minimum volumes of
pumping. Containment (physical or hydraulic) limits the
amount of uncontaminated water that requires treatment. In-
jected water can contain nutrients or electron acceptors where
bioremediation is used, or can contain enhanced oil recovery
materials (EOR) for NAPL contaminants, or can be reinfected
treated water without nutrient or EORs (U.S. EPA, 1990).
This chapter discusses pump-and-treat systems in two
categories: (1) pumping systems, and (2) treatment systems.
Pumping systems may be used for plume containment and
244
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Table 16-2. SITE Emerging Technology Program Participants with Technologies Applicable to Remediation of Contaminated
Ground Water
Developer
Atomic Energy of Canada,
Ltd. Chalk River, Ontario
(E01)
Bio-Recovery Systems, Inc.
Las Cruces, NM
(E01)
Eiectro-Pure Systems, Inc.
Amherst, NY
(E02)
Energy and Environmental
Engineering, Inc.
East Cambridge, MA
(E01)
University of Washington,
Dept. of Civil Engineering
Seattle, WA
(E02)
Wastewater Tech. Centre
Burlington, Ontario
(E02)
Technology
Chemical Treatment
Ultrafiltration
Biological Sorption
A/C Bectrocoagulation
Phase Separated and
Removal
Laser Stimulated
Photochemical Oxidation
Adsorptive Filtration
Cross-Flow Pervaporation
System
Applicable Waste
inorganic
Organic
Specific for Heavy Metals
Specific for Heavy Metals
Heavy Metals
NA
Metals
NA
NA
NA
Petroleum Byproducts, Coal-Tar
Derivatives
Non-specific
NA
Volatile Organic Compounds
NA = non applicable
Source: U.S. EPA, 1989a
Table 16-3. NATO/CCMS Projects for Remediation of Contaminated Ground Water
Treatable Contaminant
Matrix
Treatment
Organization/Site
Groundwater
Biological
Enhanced Aerobic Restoration
U.S. Air Force, Battelle
Eglin Air Force Base, FL,
United States
Chemical/Physical
Pump and Treat Groundwater
Environment Canada
Ville Mercier, Quebec
UV/Oxidation
Ultrox
San Jose, CA
United States
-------
Figure 16-1. Regions of contamination in a typical release from an underground storage tank (from Wilson et al., 1989).
plume recovery for aboveground treatment (Figures 16-5 and
16-6). Ground-water pumping systems utilize the principle
that ground water flows in response to a hydraulic gradient,
i.e., a drop in hydraulic pressure created by the combined
effects of elevation, fluid density, and gravity.
The migration of a plume away from its source area,
which is related to hydraulic containment, often can be pre-
vented by capturing the plume with a purge well. The well
must pump hard enough to overcome regional flow in the
aquifer. Hydrodynamic control of a contaminated ground-
water plume is accomplished by manipulating the hydraulic
gradient. Passive hydrodynamic controls, or interceptor sys-
tems, function by gravity. Active hydrodynamic controls rely
on injection and production wells to control the hydraulic
gradient (Canter and Knox, 1985).
Physical containment techniques include installing barri-
ers to ground-water flow (e.g., slurry walls (see Figure 16-7),
grout curtains, sheet pilings, block displacement, and clay
liners) or diverting divert uncontaminated surface water away
from waste sites or contaminated water away from clean areas
(Ehrenfield and Bass, 1984). These containment systems also
may provide for temporary containment while ground water is
removed and treated and aquifer material is decontaminated.
Contaminated ground water that is withdrawn from an
aquifer can be treated by various methods, depending on the
type(s) of contamination. Treatment methods may include one
or more of the following: (1) physical processes, such as
adsorption onto activated carbon or resins, ion exchange,
reverse osmosis, filtration, or transfer to the gaseous phase by
air stripping; (2) chemical processes, such as neutralization,
coagulation, precipitation, oxidation, or reduction reactions,
which involve inactivating or immobilizing contaminants with
chemical agents; or (3) biological processes, using conven-
tional wastewater treatment methods such as suspended growth
(e.g., activated sludge, lagoons, waste stabilization ponds, and
fluidized bed reactors) and freed film (e.g., trickling filters
and rotating biological contractors) processes (Thomas et al.,
1987).
With pumping systems that are used to bring contami-
nated ground water to the surface for treatment contaminants
are transported by advection (velocity) and dispersion. Water
velocity for pumping systems can be calculated using Darcy's
Law; however, spatial variability in hydraulic conductivity
results in a corresponding distribution of flow velocities;
therefore, contaminant removal and transport rates are distrib-
uted. Chemical contaminants in ground water also may not
move at the same rate as the water due to subsurface pro-
cesses, including sorption or retardation, ion-exchange, and
246
-------
Water Table
Depression
Oil/Water
Separation
Water Table
Depression
Pump Control
Probe Scavenger
Control Assembly
Water
Table
Depression
Pump
Perforated Well
Casing Permits
Flow of Oil and
Ground Water
Figure 162. Product recovery using two pumps in one well—a probe scavenger pump and a water table depression pump (from
Nyer, 1985).
chemical precipitation (described under soil immobilization
techniques), and bioremediation (described under soil biore-
mediation).
Monitoring water and soil cores within the plume while
pumping is occurring allows a determination of area of reme-
diation and remediation rate. These results allow rational
management of the remediation wellfield. Keely (1989) ex-
plains that, using this approach, the flow rates of extraction
wells that pump from relatively clean zones would be de-
creased, while flow rates from extraction wells that pump
from highly contaminated zones should be increased. Also,
Keely (1989) points out that the exclusive use of monitoring
points downgradient from a plume does not assist in an
understanding of plume dynamics during remediation, except
to indicate "out-of-control" conditions when contaminants are
detected.
During the continuous operation of an extraction well-
field, the level of contamination in water flowing through the
subsurface usually is decreased in a relatively short period of
time, after which a low-level residual concentration is present
in the extracted water (Figure 16-8). After the residual con-
centration in water is attained, a pump-and-treat system is
usually characterized by treatment of large volumes of slightly
contaminated water over a long period of time. In addition, if
remediation is terminated before removal of residual contami-
nation at a site, the concentration of contaminant(s) in the
aquifer water may increase due to slow release of contaminant
residuals relative to pumpage-induced water movement (Fig-
ure 16-9) (Keely, 1989). Transport processes that cause this
contaminant behavior in the subsurface include (1) diffusion
of contaminants in low-permeability sediments; (2) hydrody-
namic isolation, or dead spots, within the wellfield (3) des-
orption of contaminants from sediment surfaces; and (4)
liquid-liquid partitioning of immiscible contaminants (Keely,
1989).
One promising innovation in the use of pump-and-treat
remediation is pulsed pumping. Pulsed pumping of hydraulic
systems is the cycling of extraction or injection wells on and
off in active and resting phases (Figure 16-10) (Keely, 1989).
247
-------
Water Table
Depression
Water Table
Depression
Pump Controls
Oil/Water
Separation
Oil-Recovery
Tank
Scavenger
Control and
Pump Assembly
Perforated
Well Casing V
Water Table
Depression Pump
Water Table
Figure 16-3. Product recovery using a watertable depression pump and a floating oil/water filter (from Nyer, 1985).
The resting phase allows chemical contaminants to move
from low-permeability sediments, dead spots, sediment sur-
faces, and immiscible fluids in the subsurface into the water
phase. The pumping phase removes the minimum volume of
water at the maximum contaminant concentration. By periodi-
cally cycling selected wells, stagnation zones may be brought
into active flow paths and remediated.
When pulsed pumping systems are used, peripheral gra-
dient control must be ensured to prevent offsite migration of
contamination. If migration is slow, water would be rapidly
recovered by the high flow velocities back toward extraction
well(s) during the pumping phase. If migration is rapid, then
additional containment controls are necessary to prevent offsite
migration during the resting phase of pulsed pumping.
16.3 Biorestoration
In addition to the overviews presented by Thomas and
Ward (1989) and Lee et al. (1988), there are several milestone
publications on biological restoration of contaminated ground
waters. These publications include those that address (1)
hydrogen peroxide as a supplemental source of oxygen (Hiding
et al., 1990); (2) new approaches for site characterization,
project design, and performance evaluation (Wilson et al.,
1989); (3) methanotrophic destruction of chlorinated aliphatic
chemicals (Roberts et al., 1989); and (4) modeling aspects
(Rifai et al., 1988 and 1989).
Biological in situ treatment of subsurface contaminants in
aquifers is usually accomplished by stimulating indigenous
subsurface microorganisms to degrade organic waste con-
stituents (Thomas and Ward, 1989). The activity of microor-
ganisms is stimulated by injection of inorganic nutrients and,
if required, an appropriate electron acceptor, into aquifer
materials. Most biological in situ treatment techniques cur-
rently used are variations of techniques developed by re-
searchers at Suntech to remediate gasoline-contaminated
aquifers. The Suntech process received a patent titled Recla-
mation of Hydrocarbon Contaminated Ground Waters
248
-------
„ Groundwater Surface
Ground Surface
Oil Surface
Water Table Depression
Drainline (WTDD)
'.•V! Oil •-'•'^'•'•'.•' V. vyV.?;'. .
overy DrainJine (ORD)
Ground Surface
Sedroc/c
Underground
Tank To Treatment
Domestic
Well
• •*••»•*••*• . . . _ . * ••••
.V.V/.V.V.v. Impermeable Bedrock • •••
Ground Surface
Groundwater
Surface
,... ....t....rr^&vay'H^gTf ••••'••?•••••?•
—;» ^>^A^WQ.''.. •*& VAVAV
-------
Backhoe Keys Trench
into Bedrock
Kfill
Placed Here
Cessation
of
Pumping
(Closure?)
Time
Figure 16-7. Schematic of the preparation of a slurry wall for
physical containment of contaminated ground
water or for diversions of clean water around a
contaminated subsurface (from U.S. EPA, 1985).
Figure 18-9. Following temporary termination of pumping,
aquifer water concentration increases, or
rebounds, due to the presence of contaminant
residuals (from Keely, 1989).
0 -
Time
Figure 16-8. Decrease in aquifer water concentration caused
by pump-and-treat system where contaminant
concentration in pumped water reaches an
irreducible level that is frequently above the
regulated limit (from Keely, 1989).
sponsible for degradation of specific contaminants (Aelion et
al., 1987).
However, inoculation of a specialized microbial popula-
tion into the environment may not produce the desired degree
of degradation for a number of reasons (Goldstein et al., 1985;
Lee et al., 1988; Suflita 1989). Possible causes that may limit
the success of inoculants include both abiotic and biotic
factors. Environmental factors, such as pH, temperature, sa-
linity, and osmotic or hydrostatic pressure may act alone or
collectively to inhibit the survival of the microorganisms. The
concentration of the specific organic constituent of concern
may be too low to support growth and activity. The environ-
ment may contain substances or other organisms that are toxic
or inhibitory to the growth and activity of the inoculated
organism(s). The inoculated organism(s) may utilize some
other organic compound than the one it was selected to
metabolize. In addition, adequate mixing and transport to
ensure contact of the organism with the specific organic
constituent of concern may be difficult to achieve in ground
water. Successful inoculation of organisms into simpler, more
controllable environments (e.g., bioreactors such as wastewa-
ter treatment plants) to accomplish degradation has been
demonstrated. However, effectiveness of inoculation into un-
controlled and poorly accessible environments (e.g., the sub-
surface) is much more difficult to achieve, demonstrate, and
assess (Thomas and Ward, 1989).
In a contaminated aquifer, some regions will clean up
faster than others, and the most contaminated flow path will
be the last to be cleaned. If this flow path can be identified,
then its properties can be used to determine how much effort
and time are required to remediate the entire area. The time
required to clean the most contaminated flow path can be
determined using a modification of the relationship given by
U.S. EPA (1989c), correcting for units:
Time required to clean most contaminated flow path=
Mass of contaminant along flow path (Massc)
(Mass^Mass^^) x (Massowi/VolumewiiM) x (VolumewiiM/rime)
where
Massc/Massoxygen represents the stoichiometric amount of
oxygen required to biodegrade (mineralize) contami-
nant (hydrocarbon) present;
Massoxygen/Volumewilter represents the concentration of oxy-
gen in the ground water; and
Volumew,ter/Time represents the seepage velocity along
the contaminant flow path.
250
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Figure 16-10. Pulsed pumping removal of residual contaminant minimizes volume of water required for pumping and maximizes
contaminant concentration in pumped water (from Keely, 1989).
Generally, if the supply of mineral nutrients is adequate,
the rate of bioremediation is directly related to the rate of
supply of electron acceptor. As a result, the rate of remedia-
tion is directly proportional to the concentration of electron
acceptor in the injected water and the flow velocity of water
through the contaminated area.
When in situ bioremediation of a contaminant ground-
water plume involves using methods to enhance the process
discussed above, such as the addition of nutrients, additional
oxygen sources, or other electron acceptors, hydraulic con-
trols might be required to minimize (i.e., contain) migration of
the plume during the in situ treatment process (Thomas et al.
1987; U.S. EPA 1989c). In general, hydraulic control systems
are less costly and time consuming to install than physical
containment structures such as slurry walls. Well systems also
are more flexible, because pumping rates and well locations
can be altered as the system is operated over a period of time.
Wells should be installed under the direction of a hydrogeolo-
gist to ensure proper placement and operation.
With respect to biorestoration of aquifers, pumping-in-
jection systems can be used to (1) create stagnation (no flow)
zones at precise locations in a flow field, (2) create gradient
barriers to pollution migration, (3) control the trajectory of a
contaminant plume, and (4) intercept the trajectory of a con-
taminant plume (Schafer, 1984). The choice of a hydraulic
control method depends on geological characteristics, vari-
ability of aquifer hydraulic conductivities, background veloci-
ties, and sustainable pumping rates (Lee et al., 1988). Typical
patterns of wells that are used to provide hydraulic controls
include (1) a pair of injection-production wells, (2) a line of
downgradient pumping wells, (3) a pattern of injection-pro-
duction wells around the boundary of a plume, and (4) the
"double-cell" hydraulic containment system. The "double-
cell" system utilizes an inner cell and an outer recirculation
cell, with four cells along a line bisecting the plume in the
direction of flow (Wilson, 1984).
Well systems also serve as injection points to add materi-
als used to enhance microbial activity into the aquifer and for
control of circulation through the contaminated portion. The
system usually includes injection and production wells and
equipment for the addition and mixing of the nutrients (Lee et
al., 1988). Figure 16-11 illustrates a typical system in which
microbial nutrients are mixed with ground water and circu-
lated through the contaminated portion of the aquifer through
a series of injection and recovery wells (Raymond et al., 1978;
Thomas and Ward, 1989). Wells should be screened to ac-
commodate seasonal fluctuations in the level of the water
table and air can be supplied through a system of diffusers.
Some operational designs are closed loop in which the water
To Sewer or
Recirculate
Air
Compressor
Water Supply
Injection
*~ Well
Sparger
Clay
Figure 16-11. Typical schematic for aerobic subsurface
bioremediation (from Thomas and Ward, 1989).
251
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is recycled, thus, unused nutrients can be reinfected, disposal
of potentially hazardous ground water is avoided, and the
need for make-up water is reduced.
Materials also can be introduced into the aquifer through
the use of infiltration galleries (Figure 16-12) (Brenoel and
Brown, 1985; Thomas and Ward, 1989). Infiltration galleries
allow movement of the injection solution through the unsatur-
ated zone and the saturated zone, resulting in potential treat-
ment of source materials that may be trapped in the pore
spaces of the unsaturated zone.
Amendments to the aquifer are added to the contaminated
aquifer in alternating pulses. Inorganic nutrients are usually
added first through the injection system, followed by the
oxygen source. Simultaneous addition of the two may result in
excessive microbial growth close to the point of injection and
consequent plugging of the aquifer. High concentrations of
hydrogen peroxide (greater than 10 percent) can be used to
remove biofouling and restore the efficiency of the system.
Inorganic nutrients may be added in batch or continu-
ously, which is a more labor-intensive process. Continuous
addition of oxygen is recommended because low dissolved
oxygen levels are likely to be the rate-limiting factor in
hydrocarbon degradation. Heterogeneities in the aquifer, such
as impermeable lenses and varying hydraulic conductivities,
can hinder the distribution of nutrients and oxygen.
Both the operation and effectiveness of the system should
be monitored (Lee et al, 1988). Important operational factors
include (1) delivery of inorganic nutrients, (2) delivery of the
electron acceptor, (3) position of the delivery site in the
aquifer in relation to the contaminated portion of the plume,
and (4) effectiveness of containment and control of the con-
taminated plume.
Measurements of dissolved oxygen and nutrient levels in
ground-water samples are recommended to assess whether or
not bioremediation is successful. Increases in microbial num-
bers and/or activities in samples of aquifer materials also may
be quantified relative to (1) plume areas prior to treatment (2)
areas within the plume that did not receive treatment; and/or
(3) control areas outside the plume. Carbon dioxide levels in
ground-water samples also may be useful indicators of micro-
bial activity (Suflita, 1989).
Measurement of contaminant levels should indicate that
concentrations of contaminants are decreasing in areas that
are receiving treatment and remaining relatively unchanged in
areas that are not. If degradation pathways of specific con-
taminants are known, presence and concentrations of meta-
bolic products may be measured to determine whether or not
bioremediation is occurring. Both aquifer materials and ground-
water samples should be collected and analyzed to develop a
thorough evaluation of treatment effectiveness. The use of
appropriate control samples, e.g., assays of untreated areas or
areas outside the plume, is highly recommended to confirm
the effectiveness of the bioremediation (Suflita, 1989).
The frequency of sampling should be related to the time
expected for significant changes to occur along the most
contaminated flow path (U.S. EPA, 1989). Important consid-
erations include (1) time required for water to move from
injection wells to monitoring wells, (2) seasonal variations in
water table elevation or hydraulic gradient, (3) changes in the
concentration of dissolved oxygen or alternative electron ac-
ceptor, and (4) costs of monitoring.
Air Compressor
or Hydrogen. rz
Peroxide Tank >-• H
Nutrient Addition
Recirculated
Water and Nutrients
Figure 16-12. Use of infiltration gallery for recirculation of water
and nutrients in in situ bioremediation (from
Thomas and Ward, 1989).
16.3.1 Example of the Use of Bioremediation: A
Case Study
Lee et al. (1988) described numerous applications of the
bioremediation process used to restore contaminated aquifers.
Most applications have been in the cleanup of hydrocarbon
spills. The U.S. EPA Robert S. Kerr Environmental Research
Laboratory is presently conducting a field-scale demonstra-
tion of the use of enhanced bioremediation at the site of an
aviation fuel spill in Traverse City, Michigan. The overall
objective of the test is to provide a quantitative demonstration
of the method in order to develop a basis for process design
(Thomas and Ward, 1989).
In 1969a spill of at least 25,000 gal of aviation gasoline
from an underground storage tank at the U.S. Coast Guard Air
Station at Traverse City contaminated a shallow sandy
watertable aquifer (Figure 16-13) (Wilson et al., 1989). The
aviation gasoline was composed primarily of branched-chain
alkanes; approximately 10 percent of the spill was composed
of alkylbenzenes. Underneath the spill site a long, narrow
plume of contaminated ground water is located 15 to 17 ft
below the ground surface (corresponding closely to the sea-
sonal high and low water table at the site); it is moving in the
direction of Lake Michigan. A large contaminated plume had
moved off site, contaminating more than 40 private drinking-
252
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Figure 16-13. Former plume of contamination resulting from a spill of aviation gasoline on the U.S. Coast Guard Air Station at
Traverse City, Michigan. Beginning in 1985, the plume was intercepted by a series of wells at the Coast Guard
property boundary (from Wilson et al., 1989).
water wells. Beginning in 1985, the plume was interrupted by
a series of wells at the U.S. Coast Guard property boundary.
In 1988, the U.S. Coast Guard and the Robert S. Kerr
Environmental Research Laboratory initiated a pilot-scale in
situ bioremediation study in the area of the original spill. The
presence of alkylbenzenes is the object of regulatory concern.
Bioremediation of the site will be considered completed when
the concentration of alkylbenzenes is brought below 5 ug/L,
as specified in a consent decree between the Michigan Depart-
ment of Natural Resources and the U.S. Coast Guard.
BIOPLUME II, developed at Rice University to predict con-
taminant transport affected by oxygen-limited biodegradation
(Borden and Bedient, 1986; Borden et al., 1986; Rifai et al.,
1989b), is the model that was used to design the well system
and to estimate the time required for bioremediation.
Site characterization efforts included acquisition of cores
from the source area to determine vertical and horizontal
extent of contamination, concentration of total hydrocarbons
in the contaminated interval, and concentrations of individual
alkylbenzenes (benzene, toluene, and xylene [BTX]) (Wilson
et al., 1989). This information was used to identify the most
contaminated flow path through the spill. A series of minia-
ture monitoring wells (designated BD-7, BD-31, BD-50, BD-
62, BD-83, and BD-108) were installed along and below the
most contaminated flow path (Figure 16-14). The wells were
constructed of 3/8-in. stainless steel with a stainless steel
screen that was 6 in. long.
A set of injection wells was installed to perfuse the
contaminated area with mineral nutrients and oxygen or hy-
drogen peroxide. The nutrient solution contained 380 mg/L
ammonium chloride, 190 mg/L di-sodium phosphate, and 190
mg/L potassium phosphate. The temperature was 11 to 12°C,
and pH was near neutrality. The seepage velocity (i.e., spe-
cific discharge) of the injected water in the aquifer averaged
10 ft per day. A tracer test using chloride was conducted for
each monitoring well to determine the actual seepage velocity
along the flow path to that particular well. Typical data
obtained from tracer testing are given in Figure 16-15.
Injection began the first week of March 1988. The system
was first acclimated with pure oxygen and then switched to
perfusion with hydrogen peroxide. The concentration of hy-
drogen peroxide was increased slowly to allow time for
microbial acclimation to concentrations of hydrogen perox-
ide, which are generally toxic to most heterotrophic bacteria.
253
-------
Elevation in
Feet Above MSL
Injection
Wells
BD'50 BD'62
B°'83
Figure 16-14. Cross section of a demonstration project for bioremediation of the aviation gasoline spill at Traverse City, Michigan
(from Wilson et al., 1989).
200
150
roo
§
50 .
Chloride Breakthrough
BD 50B-2
PJ n D n gap
40
80
120 160
Time: Hours
200
—i 1 1
240 280
Figure 16-15. Tracer test in the flow line between the injection wells and a miniature monitoring well (BD 50B-2) 50 feet away (from
Wilson et al., 1989).
254
-------
The schedule of application of oxygen or hydrogen peroxide
is shown in Figure 16-16.
Table 16-4. Stoiohiometry of Aerobic Bioremediation of the
Aviation Fuel Spill
Monitoring Wells
Oxygen required
BD 31-2
BD 506-2
400
50 100 150 200 250 300 350
Julian Date
Figure 16-16. Schedule of application of oxygen or hydrogen
peroxide in the first year of the demonstration
project (1988) (from Wilson et al., 1989).
The concentration of fuel hydrocarbons in the most-
contaminated flow path averaged 7,500 mg/kg of aquifer
material. Based on the concentration of hydrocarbons, the
length of the contaminated portion of the flow path, and an
assumed stoichiometry for microbial respiration, the total
oxygen (Table 16-4) required to remediate the flow paths to
the monitoring wells at 31 and 50 ft was estimated.
The interval between the injection wells and the monitor-
ing wells was considered remediated when detectable oxygen
broke through and alkylbenzenes disappeared. The interval to
the monitoring well at 31 ft was remediated after 220 days
(Julian Date 281) (Figure 16-17). The interval to the monitor-
ing well at 50 ft was remediated after 270 days (Julian Date
331) (Figure 16-18).
The seepage velocity of the ground water (as determined
by the chloride tracer tests) was multiplied by the concentra-
tion of oxygen or hydrogen peroxide in the injection wells to
determine the instantaneous flux of oxygen or hydrogen per-
oxide along the flow path. The cumulative flux at the time of
remediation was considered the actual oxygen demand for
remediation (Table 16-5).
The aquifer was purged of alkylbenzenes very quickly.
The aviation gasoline was composed primarily of branched-
chain alkanes, while only about 10 percent of the original spill
was composed of alkylbenzenes. The quantity of oxygen and
mg Obiter pore water
Estimated based on:
Total Fuel Hydrocarbons
BTX only (8/87)
BTX only (3/88)
Actually required
62,212
8,710
2,364
2,989
90,000
12,000
3,420
2,952
Source: Wilson et al., 1989
hydrogen peroxide required to remove alkylbenzenes from
the wells agreed closely with the projected demand of the
alkylbenzenes alone.
The flow paths to the monitoring wells at 31 and 50 ft
from the injection wells were remediated when only a small
fraction of the total oxygen demand of the spill had been
supplied. Some of the alkylbenzenes may have been washed
from the source area by the flow of water, because of their
relatively high water solubility. The significance of transport
in the aqueous phase was evaluated by comparing the retarda-
tion factor of each alkylbenzene in the most contaminated
interval to the number of pore volumes that had been deliv-
ered to a particular point. The results of this evaluation
indicated that benzene easily could have been removed by
water transport and a fraction of the toluene may have been
removed by this process. The removal of xylenes, ethylbenzene,
however, or trimethylbenzene would not be expected.
An additional portion of the alkylbenzenes may have
been removed by anaerobic processes before the front of
oxygen passed through. Water from anaerobic regions of the
demonstration area contained significant concentrations of
volatile fatty acids and was visibly turbid with microorgan-
isms.
The spill was cored in August 1987 to provide informa-
tion to design the pilot-scale field study and cored again in
March 1988, just prior to the initiation of the study, to define
the initial conditions (Table 16-5). In general, the concentra-
tion of alkylbenzenes declined from 1987 to 1988. Some of
the alkylbenzenes may have been removed from the source
area after the sampling in 1987 and before the initiation of the
remediation action in 1988. This removal was probably due to
anaerobic biological processes.
After 8 months of perfusion of the aquifer with mineral
nutrients and oxygen sources, results of analyses of core
samples taken at 31 ft from the injection wells showed that the
aliphatic hydrocarbons remained near their initial concentra-
tions, while the concentrations of the alkylbenzenes had de-
creased to below analytical detection limits. Since only a
minor fraction of their oxygen demand had been supplied
when the alkylbenzenes had disappeared from the aquifer, it
255
-------
20
15 -i
fc 10
o
I
.0— DO
•*— BTX
600
. 500
P 400
300
- 200
. 100
co
.o
100
200
Julian Date
300
Figure 16-17. Breakthrough of oxygen and depletion of alkylbenzenes (BTX) in a miniature monitoring wail (BD 31-2) 31 feet from
the injection wells (from Wilson et al., 1989).
600
200
Julian Date
300
Figure 16-18. Breakthrough of oxygen and depletion of alkylbenzenes (BTX) in a miniature monitoring well (BD 50B-2) 50 feet from
the injection wells (from Wilson et al., 1989).
256
-------
Table 16-5. Changes in Concentrations of Alkylbenzenes and Total Fuel Hydrocarbon in Core Material During Bioremediation of
an Aquifer Contaminated with Aviation Gasoline
Date Oil and Grease
Core
Number
Fuel Hydrocarbon Benzene Toluene
mg/kg wet sample
Ethylbenzene
Xylenes
Background conditions in an unweathered part of the spill area., June 1988.
50R6 12,150 1.0 107
50R7 5,220 1.0 170
57
24
Preliminary sampling used to design the bioremediation project near monitoring well BD-31-2, August 1987.
50A3 4,310 5,590 0.6 235 33
50114 4,130 6,500 0.3 444 12
50D18 1,130 2,500" 0.7 112 11
Sampled after four months of perfusion with mineral nutrients and oxygen, June 1988.
50T3 3,330" 1.4 f
50W3 4,800" 1.5 f
7.3
13
Sampled after eight months of perfusion with mineral nutrients and oxygen, October 1988.
50AE4 8,400 <0.3 <0.3 <0.3
50AE5 2,370" <0.3 <0.3 <0.3
218
100
121
48
39
23
41
<0.3
<0.3
"These cores included some uncontaminated material.
Wilson et al., 7989
appears that the nonaromatic fraction of the spill remained in
the aquifer.
When the region at 31 ft from the injection wells was
cored in March 1989, almost all of the petroleum hydrocar-
bons had been removed, including the branched-chain al-
kanes. The study is continuing at the site, and additional
information may be obtained from the U.S. EPA Robert S.
Kerr Environmental Research Laboratory.
16.3.2 Advantages and Limitations in the Use of
In Situ Bioremediation
In situ bioremediation has been used most often, and with
reasonably good success, to treat gasoline spills. It has been
combined with other treatment processes to reduce organic
contaminants in aquifers. In most cases, contaminated ground
water is withdrawn, heated by a physical, chemical, or bio-
logical aboveground treatment technique, and then recharged
to the aquifer after aeration and addition of nutrients. The role
of bioremediation in such combination treatment schemes
often is difficult to assess.
There are a number of advantages and disadvantages in
the use of in situ bioremediation (Lee et al., 1988). Unlike
other aquifer remediation technologies, it often can be used to
treat contaminants that are sorbed to aquifer materials or
trapped in pore spaces. In addition to treatment of the satu-
rated zone, organic contaminants held in the unsaturated and
capillary zones can be treated when an infiltration gallery is
used. Complete aerobic biodegradation (mineralization) of
organic compounds usually produces carbon dioxide, water,
and an increase in cell mass.
The time required to treat subsurface contamination using
in situ bioremediation often can be faster than withdrawal and
treatment processes. A gasoline spill was remediated in 18
months using in situ bioremediation; pump-and-treat tech-
niques were estimated to require 100 years to reduce the
concentrations of gasoline to potable water levels (Raymond
et al., 1976). Also, in situ bioremediation often costs less than
other remedial options. The areal zone of treatment using
bioremediation can be larger than with other remedial tech-
nologies because the treatment moves with the plume and can
reach areas that would otherwise be inaccessible.
There are also disadvantages to in situ bioremediation
programs (Lee et al., 1988). Many organic compounds in the
subsurface are resistant to degradation. In situ bioremediation
usually requires an acclimated population of microorganisms;
however, an acclimated population may not have developed
for recent spills or for recalcitrant compounds. Heavy metals
and toxic concentrations of organic compounds may inhibit
activity of indigenous microorganisms. Injection wells may
become clogged from profuse microbial growth resulting
from the addition of nutrients and oxygen. Nutrients added to
the aquifer must be contained within the treatment zone
because their transport to surface waters could result in eu-
trophication. Additionally, using nitrate as an electron accep-
tor may result in unacceptable levels of unused nitrate being
transported through the ground water to potable ground-water
or surface water supplies.
Metabolites resulting from partial degradation of organic
contaminants also may impart objectionable tastes and odors.
For example, incomplete degradation of gasoline under low
dissolved oxygen conditions has been shown to result in
phenol production, and phenol degradation required more
aerobic conditions (Raymond et al., 1978). Increased micro
bial biomass can exert an oxygen demand that can form
anaerobic conditions in the aquifer, which may result in the
production of hydrogen sulfide or other objectionable by-
products.
257
-------
In situ bioremediation is difficult to implement in low-
permeability aquifers that do not permit the transport of
adequate supplies of nutrients or oxygen to active microbial
populations. In addition, bioremediation projects require con-
tinuous monitoring and maintenance for successful treatment.
Costs associated with in situ bioremediation include (1)
site characterization, (2) remedial design, (3) system design,
(4) system installation, (5) materials and operating expenses,
and (6) monitoring (U.S. EPA, 1989c).
16.4 References
Aelion, C.M., C.M. Swindell, andF.K. Pfaender. 1987. Adap-
tation to and Biodegradation of Xenobiotic Compounds
by Microbial Communities from a Pristine Aquifer. Appl.
Environ. Microbiol. 53:2212-2217.
Borden, R.C. and P.B. Bedient. 1986. Transport of Dissolved
Hydrocarbons Influenced by Reaeration and Oxygen Lim-
ited Biodegradation. I. Theoretical Development. Water
Resources Research 22:1973-1982.
Borden, RC., P.B. Bedient, M.D. Lee, C.H. Ward, and J.T.
Wilson. 1986. Transport of Dissolved Hydrocarbons In-
fluenced by Reaeration and Oxygen Limited Biodegrada-
tion. II. Field Application. Water Resources Research
22:1983-1990.
Brenoel, M. and R.A. Brown. 1985. Remediation of a Leaking
Underground Storage Tank with Enhanced Bioreclama-
tion. In: Proc. Fifth Nat. Symp. on Aquifer Restoration
and Ground Water Monitoring, National Water Well
Association, Dublin, OH, pp. 527.
Canter, L.W. and R.C. Knox. 1985. Ground Water Pollution
Control. Lewis Publishers, Chelsea, MI.
Ehrenfield, J. and J. Bass. 1984. Evaluation of Remedial
Action Unit Operations of Hazardous Waste Disposal
Sites. Pollution Technology Review No. 110, Noyes Pub-
lications, Park Ridge, NJ.
Goldstein, R.M., L.M. Mallory, and M. Alexander. 1985.
Reasons for Possible Failure of Inoculation to Enhance
Biodegradation. Appl. Environ. Microbiol. 50:977-983.
Ruling, S.G., B.E. Bledsoe, and M.V. White. 1990. Enhanced
Bioremediation Utilizing Hydrogen Peroxide as a Supple-
mental Source of Oxygen: A Laboratory and Field Study.
EPA/600/2-90/006 (NTIS PB90-183435/AS)
Keely, J.F. 1989. Performance Evaluations of Pump-and-
Treat Remediations. Superfund Groundwater Issue Paper
No. 5. EPA/540/4-89/005.
Lee, M.D. and C.H. Ward. 1986. Ground Water Restoration.
Report submitted to JACA Corporation, Fort Washing-
ton, PA.
Lee, M.D., J.M. Thomas, RC. Borden, P.B. Bedient, J.T.
Wilson, and C.H. Ward. 1988. Biorestoration of Aquifers
Contaminated with Organic Compounds. CRC Critical
Reviews In Environmental Control 18:29-89.
Nyer, E.K. 1985. Groundwater Treatment Technology. Van
Nostrand Reinhold, New York, NY.
Palmer, C.D. and RL. Johnson. 1989. Physical processes
Controlling the Transport of Non-Aqueous Phase Liquids
in the Subsurface. In: Transport and Fate of Contami-
nants in the Subsurface. EPA/625/4-89/019, Chapter 3.
Raymond, R.L. 1974. Reclamation of Hydrocarbon Contami-
nated Ground Water. U.S. Patent 3,846,290, Nov. 5.
Raymond, R.L., V.W. Jamison, and J.O. Hudson. 1976. Ben-
eficial Stimulation of Bacterial Activity in Groundwaters
Containing Petroleum Products. AIChE Symposium Se-
ries 73:390.
Raymond, R.L., V.W. Jamison, J.O. Hudson, RE. Mitchell,
and V.E. Farmer. 1978. Field Application of Subsurface
Biodegradation of Gasoline in a Sand Formation. API
Publication 4430, American Petroleum Institute, Wash-
ington, DC.
Rifai, H. S., P.B. Bedient, J.T. Wilson, K.M. Miller, and J.M.
Armstrong. 1988. Biodegradation Modeling at Aviation
Fuel Spill Site. J. Environ. Engineering 114(5)1007-
1029.
Rifai, H.S., P.B Bedient, RC. Borden, and J.F. Haasbeek.
1989. BIOPLUME II-Computer Model of Two-Dimen-
sional Contaminant Transport Under the Influence of
Oxygen Limited Biodegradation in Ground Water (User's
Manual Version 1.0; Preprocessor Source Code Version
1.0; Source Code Version 1.0). EPA/600/8-88/093 (NTIS
PB89-151120/AS).
Roberts, P. V., L. Semprini, G.D. Hopkins, D. Grbic-Galic,
P.L. McCarty, and M. Reinhard. 1989. In-Situ Aquifer
Restoration of Chlorinated Aliphatics by Methanotrophic
Bacteria. EPA/60012-89/033 (NTIS PB89-219992).
Sale, T, and K. Piontek. 1989. In Situ Removal of Waste
Wood-Treating Oils from Subsurface Materials. Presented
at Forum on Remediation of Wood Preserving Sites,
Technical Assistance to U.S. EPA Region IX (Edwin
Earth, U.S. EPA, Cincinnati, OH, and John Matthews,
U.S. EPA, Ada, OK, Technical Coordinators).
Schafer, J.M. 1984. Determining Optimum Pumping Rates
for Creation of Hydraulic Barriers to Ground Water Pol-
lutant Migration. In: Proc. Fourth Nat. Symp. on Aquifer
Restoration and Ground Water Monitoring, National Water
Well Association, Dublin, OH, pp. 50-62.
Suflita, J.M. 1989. Microbiological Principles Influencing the
Restoration of Aquifers. In: Transport and Fate of Con-
taminants in the Subsurface, EPA/625/4-89/0 19, Chapter
258
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Thomas, J.M. and C.H. Ward. 1989. In Situ Biorestoration of
Organic Contaminants in the Subsurface. Environ. Sci.
Technol. 23:760-766.
Thomas, J.M., M.D. Lee, P.B. Bedient, R.C. Borden, L.W.
Canter, and C.H. Ward. 1987. Leaking Underground
Storage Tank Remediation with Emphasis on In Situ
Biorestoration. EPA/600/2-87/108 (NTIS PB87-168084/
REB).
U.S. Environmental Protection Agency (EPA). 1985. Reme-
dial Action at Waste Disposal Sites. EPA/625/6-85/006
(NTIS PB87-201034).
U.S. Environmental Protection Agency (EPA). 1989a. The
Superfund Innovative Technology Evaluation Program:
Technology Profiles. EPA/540/5-89/013.
U.S. Environmental Protection Agency (EPA). 1989b. Dem-
onstration of Remedial Action Technologies for Con-
taminated Land and Groundwater. In: proceedings and
Appendices, Third International Conference, NATO Com-
mittee on Challenges of Modem Society (CCMS)
(Montreal, Canada).
U.S. Environmental Protection Agency (EPA). 1989c. Biore-
mediation of Hazardous Waste Sites Workshop: Speaker
Slide Copy and Supporting Information. EPA/CERI-89-
11 (NTIS PB89-169205/REB).
US. Environmental Protection Agency (EPA). 1990. Basics of
Pump-and-Treat Ground-Water Remediation Technology.
EPA/600113-90/003.
Wilson, J.L. 1984. Double-Cell Hydraulic Containment of
Pollutant Plumes. In: Proc. Fourth Nat. Symp. on Aquifer
Restoration and Ground Water Monitoring, National Water
Well Association, Dublin, OH, pp. 65-70.
Wilson, J.T., L.E. Leach, J. Michalowski, S. Vandegrift, and
R. Callaway. 1989. In Situ Bioremediation of Spills from
Underground Storage Tanks: New Approaches for Site
Characterization, Project Design, and Evaluation of Per-
formance. EPA/600/2-89/042 (NTIS PB89-219976/AS).
• U.S. GOVERNMENT PRINTING OFFICE:1994 -550-001/00182
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