vxEPA
United States
Environmental Protection
Agency
Office of Research and
Development
Washington DC 20460
EPA/625/6-89/025a
June 1990
Assessing the
Geochemicai Fate of
Deep-Well-lnjected
Hazardous Waste
A Reference Guide
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EPA/625/6-89/025a
June 1990
Assessing the Geochemical Fate of
Deep-Well-lnjected Hazardous Waste:
A Reference Guide
U.S. Environmental Protection Agency
Office of Research and Development
Center for Environmental Research Information
Cincinnati, OH 45268
Robert S. Kerr Environmental Research Laboratory
Ada, Oklahoma 74820
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Notice
This document has been reviewed in accordance with the U.S. Environmental Protection Agency's
peer and administrative review policies and approved for publication. Mention of trade names of
commercial products does not constitute endorsement or recommendation for use.
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Acknowledgments
Although many individuals contributed to the preparation of this document, the assistance of the
individuals listed below is especially acknowledged.
Major Author:
J. Russell Boulding, Eastern Research Group, inc. (ERG), Arlington, Massachusetts
ERG Project Management and Technical Writing/Editing:
Anne C. Jones, Eastern Research Group, Inc., Arlington, Massachusetts
EPA Project Management:
Carol Grove, EPA CERI, Cincinnati, Ohio
Jerry Thornhill, EPA/ORD, Robert S. Kerr Environmental Research Laboratory, Ada, Oklahoma
Reviewers:
Robert E. Smith, EPA Office of Drinking Water/UIC Branch, Washington, D.C.
Dr. William Roy, Illinois State Geological Survey, Champaign, Illinois
in
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Contents
Chapter
1 OVERVIEW OF DEEP-WELL INJECTION OF
HAZARDOUS WASTES IN THE UNITED STATES ...................... 1
1 .1 Identifying Hazardous Wastes ............................. 1
1.1.1 Toxicity ..................................... 1
1.1.2 Reactivity .................................... 1
1.1.3 Corrosivity ................................... 2
1.1.4 Ignitability .................................... 2
1 .2 Sources, Amounts and Composition of Deep-Well-
Injected Wastes ..................................... 2
1 .3 Geographic Distribution of Hazardous Waste
Injection Wells ..................................... 5
1 .4 Design and Construction of Deep-Injection Wells ................... 5
1.4.1 Surface Equipment Used in Waste Disposal .................. 5
1 .4.2 Injection-Well Construction ........................... 8
2 PROCESSES AFFECTING THE GEOCHEMICAL FATE
OF DEEP-WELL-INJECTED WASTES ............................ 11
2.1 Overview of Fate-Influencing Processes in Chemical
Systems ......................................... 11
2.1.1 Key Characteristics of Chemical Systems ................... 11
2.1 .2 Fate-Influencing Processes in the Deep-Well
Environment .................................. 13
2.2 Partition Processes ................................... 15
2.2.1 Acid-Base Reactions .............................. 16
2.2.2 Adsorption and Desorption ........................... 16
2.2.3 Precipitation and Dissolution .......................... 20
2.2.4 Immiscible-Phase Separation ......................... 21
2.3 Transformation Processes ............................... 21
2.3.1 Neutralization .................................. 21
2.3.2 Complexation .................................. 23
2.3.3 Hydrolysis .................................... 24
2.3.4 Oxidation-Reduction .............................. 25
2.3.5 Catalysis .................................... 27
2.3.6 Polymerization ................................. 27
2.3.7 Thermal Degradation .............................. 28
2.3.8 Biodegradation ................................. 28
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Table of Contents (Continued)
Chapter Page
2.4 Transport Processes 29
2.4.1 Hydrodynamic Dispersion 29
2.4.2 Osmotic Potential 30
2.4.3 Particle Migration 31
2.4.4 Density/Viscosity Differences 31
2.5 Interaction of Partition, Transformation, and
Transport Processes 31
3 MAJOR ENVIRONMENTAL FACTORS AFFECTING DEEP-WELL-INJECTION
GEOCHEMICAL PROCESSES 37
3.1 Major Environmental Factors Influencing Geochemical-
Fate Processes 37
3.1.1 pH '.'.'.'.'.'. '.'.'.'.'.'. '.'.37
3.1.2 Eh and Other Redox Indicators 39
3.1.3 Salinity and Specific Conductance 40
3.1.4 Reservoir Matrix 40
3.1,5 Temperature and Pressure 44
3.2 Geochemical Characteristics of Deep-Well
Injection Zones 48
3.2.1 Lithology i 48
3.2.2 Brine Chemistry 49
3.3 Influence of Environmental Factors on Waste/
Reservoir Compatibility 49
3.3.1 Well Plugging 56
3.3.2 Well-Casing and Confining-Formation Failure 59
3.3.3 Well Blowout 59
3.4 Influence of the Deep-Well Environment on
Biodegradation 59
3.4.1 Occurrence of Microbes 59
3.4.2 Degradation of Organic Compounds in
Anaerobic Conditions 61
3.4.3 Microbial Ecology 64
4 GEOCHEMICAL CHARACTERISTICS OF HAZARDOUS WASTES 77
4.1 Inorganic vs. Organic Hazardous Wastes 77
4.2 Chemical Properties of Inorganic Hazardous Wastes 78
4.2.1 Major Processes and Environmental Factors
Affecting Geochemical Fate of Hazardous
Inorganics 78
4.2.2 Known Properties of Listed Hazardous Inorganics 81
4.3 Chemical Properties of Organic Hazardous Wastes 81
4.3.1 Halogenated Aliphatic Hydrocarbons 84
4.3.2 Halogenated Ethers 86
4.3.3 Monocyclic Aromatic Hydrocarbons and Halides 86
4.3.4 Phthalate Esters 88
4.3.5 Polycyclic Aromatic Hydrocarbons 88
4.3.6 Nitrogenous Compounds 89
4.3.7 Pesticides 90
VI
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Table of Contents (Continued)
Chapter Page
4.4 Locating Data on Specific Hazardous Substances 90
4.4.1 Basic References 91
4.4.2 Sources of Information on Geochemical Fate 92
4.4.3 Computerized Databases 93
4.4.4 Benchmark and Structure-Activity Concepts 93
5 METHODS AND MODELS FOR PREDICTING THE GEOCHEMICAL FATE
OF DEEP-WELL-INJECTED WASTE 97
5.1 Basic Approaches to Geochemical Modeling 97
5.1.1 Model Evaluation 97
5.1.2 Model Deficiencies . . . 97
5.2 Specific Methods and Models 98
5.2.1 Aqueous-and Solution-Geochemistry Computer Codes 99
5.2.2 Adsorption 100
5.2.3 Biodegradation 106
5.2.4 Hydrolysis 109
5.2.5 Chemical Transport 110
6 FIELD SAMPLING AND LABORATORY PROCEDURES
AND PROTOCOLS 119
6.1 Overview 119
6.1.1 Chapter Organization 119
6.1.2 Selecting Sampling Methods and Laboratory
Procedures 119
6.2 Waste/Reservoir Characterization 120
6.2.1 The Waste Stream 120
6.2.2 Reservoir Lithology 121
6.2.3 Formation Water 122
6.2.4 Microbiology 126
6.3 Waste/Reservoir Interaction Tests 126
6.4 Geochemical Processes 128
6.4.1 Adsorption Isotherms 128
6.4.2 Hydrolysis 128
6.4.3 Biodegradation 130
6.5 Quality-Assurance/Control Procedures 130
6.6 Annotated Bibliography 131
6.6.1 How to Use this Bibliography 131
6.6.2 Annotations 131
7 CASE STUDIES OF DEEP-WELL INJECTION OF
INDUSTRIAL WASTE 141
7.1 Use of Field Studies in Geochemical Fate Assessment 141
7.1.1 Monitoring Wells 141
7.1.2 Backf lush ing of Injected Wastes 143
7.2 Case Study No. 1: Pensacola, Florida (Monsanto) 143
7.2.1 Injection-Facility Overview 143
7.2.2 Injection/Confining-Zone Lithology and Chemistry 145
7.2.3 Chemical Processes Observed 145
VII
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Table of Contents (Continued)
Chapter Page
7.3 Case Study No. 2: Pensacola, Florida (American Cyanamid) 147
7.3.1 Injection-Facility Overview 147
7.3.2 Injection/Confining-Zone Lithology and Chemistry 148
7.3.3 Chemical Processes Observed 148
7.4 Case Study No. 3: Belle Glade, South Central Florida 148
7.4.1 Injection-Facility Overview 148
7.4.2 Injection/Confining-Zone Lithology and Chemistry 151
7.4.3 Chemical Processes Observed 151
7.5 Case Study No. 4: Wilmington, North Carolina 152
7.5.1 Injection-Facility Overview 152
7.5.2 Injection/Confining-Zone Lithology and Chemistry 153
7.5.3 Chemical Processes Observed 153
7.6 Case Study No. 5: Illinois Hydrochloric Acid Injection Well 154
7.6.1 Injection-Facility Overview 154
7.6.2 Injection/Confining-Zone Lithology and Chemistry 155
7.6.3 Chemical Processes Observed 155
7.7 Case Study No. 6: Texas Petrochemical Plant 155
7.7.1 Injection-Facility Overview 155
7.7.2 Injection/Confining-Zone Lithology and Chemistry 155
7.7.3 Chemical Processes Observed 155
APPENDIX A. Section and Table Index for EPA Priority Pollutants 161
APPENDIX B. Ground-Water Contaminant Fate (Adsorption and
Biodegradation) and Transport Studies Indexed by
Organic Compound 167
VIII
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List of Tables
Table
Page
Table 1-1 Typical Physical/Chemical Properties of Hazardous
Components in Deep-Weil-Injected Hazardous Wastes 2
Table 1 -2 Estimated Volume of Deep-Well-Injected Wastes by
Industrial Category, 1983 3
Table 1 -3 Waste Characteristics of 108 Hazardous Waste Wells
Active in 1983 in the United States 4
Table 1 -4 Historical Trends in Distribution of Industrial-
Waste Injection Wells 6
Table 1 -5 Applicability of Tests That May Be Used for Mechanical
Integrity Verification 9
Table 2-1 Characteristics of Chemical Processes That May Be
Significant in the Deep-Well Environment 12
Table 2-2 Near-Surface Geochemical Processes and Their Relevance
to the Deep-Well Environment 14
Table 2-3 Significance of Chemical Processes in the Deep-Well
Environment 15
Table 2-4 Acid-Base Characteristics of Toxic Organics 16
Table 2-5 Major Intermolecular Interactions Involved in Adsorption
in the Deep-Well Environment 19
Table 2-6 Examples of the Effects of Transformation Processes on
the Toxicity of Substances 22
Table 2-7 Listed Hazardous Organic Wastes for which Hydrolysis May
Be a Significant Transformation Process in the Deep-Well
Environment 24
Table 2-8 Amenability of Organic Functional Groups to Hydrolysis 25
Table 2-9 Redox Reactions in a Closed Ground-Water System 26
Table 2-10 Relative Oxidation States of Organic Functional
Groups 27
Table 2-11 Susceptibility of Organic Compounds to Oxidation in Water 27
Table 2-12 Summary Descriptions of the Major Types of Biological
Transformation Processes 29
Table 2-13 Physical Parameters Affecting Particle
Migration in Porous-Media Flow 32
Table 3-1 Effects of pH on Deep-Well Geochemical Processes and
Other Environmental Factors 38
Table 3-2 Important Characteristics of Silicate Clay Minerals 42
Table 3-3 Mineral Composition and Particle-Size Distribution of
Core Samples of Upper Frio Formation, Texas 44
Table 3-4 Effect of Particle Size on Cation-Exchange Capacity (CEC)
of Natural Streambed Sediments, San Mateo County, California 45
Table 3-5 Temperature and Pressure at Different Depths 46
Table 3-6 Effects of Increased Temperature and Pressure on
Waste-Rock Mixtures 47
Table 3-7 Lithology and Age of Geologic Formations Used for
Injection of Industrial Wastes 48
Table 3-8 U.S. Geologic Formations Being Used for Hazardous Waste
Disposal 50
IX
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List of Tables (Continued)
Table
Page
Table 3-9 Sources of National, Regional, and State Information on
Suitability of Geologic Formations for Deep-Well Injection 52
Table 3-10 Implications of Mechanisms for Brine Formation on
Movement of Injected Wastes 53
Table 3-11 Selected Parameters of Brines from Formations Used for
Deep-Well Injection of Hazardous Wastes 55
Table 3-12 Processes Significant in Different Types of Waste-
Reservoir Interactions 55
Table 3-13 Causes of Well Plugging and Possible Remedial Actions 56
Table 3-14 Effectiveness of Various Metal Ions in Controlling
Formation-Water Sensitivity 57
Table 3-15 Examples of Waste/Reservoir Incompatibility 58
Table 3-16 Organic Compounds Degraded Under Denitrifying
Conditions 62
Table 3-17 Organic Compounds Degraded Under Sulfate-Reducing
Conditions 62
Table 3-18 Organic Compounds Degraded Under Methanogenic
Conditions 63
Table 4-1 Inorganic Hazardous Wastes (Excluding Radioactive
Elements) 79
Table 4-2 Major Processes and Environmental Factors Affecting the
Geochemical Fate of Inorganic Hazardous Wastes 80
Table 4-3 Geochemical Properties of Listed Metals and Nonmetals 80
Table 4-4 Major Processes and Environmental Factors Affecting the
Geochemical Fate of Organic Hazardous Wastes 84
Table 4-5 Geochemical Processes Affecting the Fate of Halogenated
Aliphatic Hydrocarbons 85
Table 4-6 Geochemical Processes Affecting the Fate of Halogenated
Ethers 86
Table 4-7 Geochemical Processes Affecting the Fate of Monocyclic
Aromatic Hydrocarbons and Halides 87
Table 4-8 Geochemical Processes Affecting the Fate of Phthalate
Esters 88
Table 4-9 Geochemical Processes Affecting the Fate of Polycyclic
Aromatic Hydrocarbons (PAHs) 89
Table 4-10 Geochemical Processes Affecting the Fate of Nitrogenous
and Miscellaneous Compounds 90
Table 4-11 Geochemical Processes Affecting the Fate of Pesticides 91
Table 5-1 Definitions of Terms Used in Chemical-Fate Modeling 98
Table 5-2 Aqueous- and Solution-Geochemistry Models of Potential
Value for Modeling Deep-Well Injection 101
Table 5-3 Applicability of Methods and Models for Predicting
Adsorption in the Deep-Well Environment 102
Table 5-4 Results of Adsorption Experiments with Organic Compounds
at Simulated Deep-Well Conditions 105
Table 5-5 Redox Zones for Biodegradation of Organic Micropollutants 109
Table 5-6 Integrated Ground-Water Chemical-Transport Models 111
Table 5-7 Two-Step Ground-Water Chemical-Transport Models 111
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List of Tables (Continued)
Table Page
Table 6-1 Basic Parameters for Characterizing Wastewater 121
Table 6-2 Physical Properties of Reservoirs Important in Deep-Well
Geochemical Fate Assessment 122
Table 6-3 Chemical Analysis Methods for Reservoir Rock 123
Table 6-4 Chemical Properties of Reservoir Fluids Important in
Deep-Well Geochemical Fate Assessment 123
Table 6-5 Classification of Dissolved Species in Deep-Well
Formation Water 124
Table 6-6 Chemical Constituents of Formation Waters Analyzed in
Studies Related to Deep-Well Injection 125
Table 6-7 Methods for Subsurface Microbial Characterization 127
Table 6-8 Summary of Waste-Reservoir Compatibility/Interaction
Studies 129
Table 6-9 Topical Index to Annotated Bibliography 132
Table 7-1 Summary of Case Studies 142
XI
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List of Figures
Figure
Page
Figure 1-1 Regulatory Status of Class I Wells in the United States 7
Figure 1 -2 Typical Above-Ground Components of a Subsurface Waste
Disposal System 8
Figure 2-1 Hypothetical Model for Population Changes and Metabolism
of a Chemical Modified by Mineralizing and Co-metabolizing
Populations 30
Figure 2-2 Effects of Dispersion, Adsorption, and Biodegradation
on the Time Change in Concentration of an Organic Compound
in an Aquifer Observation Well 33
Figure 3-1 Geologic Features Significant in Deep-Waste Injection
Well Site Evaluation, and Locations of Industrial-Waste
Injection Systems, 1966 53
Figure 3-2 Site Suitability for Deep-Well Injection of Industrial
Waste, and Locations of Industrial Waste Disposal Wells,
1976 54
Figure 4-1 Periodic Chart of the Elements, Showing Position of
Toxic Metals and Nonmetals 81
Figure 4-2 Types of Metal Species in Water 82
Figure 4-3 Distribution of Molecular and Ionic Species of Divalent
Cadmium at Different pH Values 82
Figure 4-4 Distribution of Molecular and Ionic Species of Divalent
Lead at Different pH Values 82
Figure 4-5 Distribution of Molecular and Ionic Species of Divalent
Mercury at Different pH Values 83
Figure 5-1 Relative Trade-offs Between Physical (Microcosm) and
Mathematical Models as Affected by Effluent Complexity 99
Figure 5-2 Freundlich Isotherm for Phenol Adsorbed on Frio Core 104
Figure 5-3 Proposed Geochemical Model of Waste after Injection
into the Subsurface 108
Figure 7-1 Location of Three Monitoring and Two Injection Wells,
Monsanto Facility 144
Figure 7-2 Hydrology of the Lower Limestone of Floridan Aquifer in
Northwest Florida 145
Figure 7-3 Generalized North-South Geologic Section Through Southern
Alabama and Northwestern Florida 146
Figure 7-4 Monsanto Injection Facility Hydrogeologic Cross-Section 147
Figure 7-5 Location of the American Cyanamid Injection Site and
Monitoring Wells 149
Figure 7-6 American Cyanamid Injection Facility Hydrogeologic Cross
Section 150
Figure 7-7 Index Map of Belle Glade Area and Potentiometric-Surface
Map of Floridan Aquifer in South Florida 151
Figure 7-8 Generalized Hydrogeologic Section between Belle Glade
and the Straits of Florida 152
Figure 7-9 Map of Wilmington, North Carolina Waste-Injection and
Observation Wells 153
Figure 7-10 Diagram Showing Construction Features and Lithologic Log
of Well 14, Wilmington, North Carolina 154
XIII
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List of Figures (Continued)
Figure Page
Figure 7-11 North-South Cross Section Showing Oil Wells and
Inclination of Major Formations, Texas Petrochemical Plant 156
Figure 7-12 Texas Petrochemical Plant Injection Well ! . ! 156
XIV
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PREFACE
The geochemical fate of deep-well-injected wastes
must be thoroughly understood to help avoid problems
when incompatibility between the injected wastes and
the injection-zone formation is a possibility. An under-
standing of geochemical fate also will be useful when a
geochemical no-migration demonstration must be
made. This reference guide was written to address
both of these needs by presenting state-of-the-art infor-
mation on the geochemical fate of hazardous deep-
well-injected wastes. Furthermore, operators of any
new industrial-waste injection well who must consider
the possibility of incompatibility will find this guide help-
ful in identifying geochemical reactions of potential con-
cern and methods for testing incompatibility.
U.S. EPA regulations (53 Federal Register 28118-
28157, July 26,1988) stipulate that deep-well injection
of hazardous wastes is allowed only if either of two no-
migration standards is met (40 CFR 148.20[a][1]):
1) Fluid movement conditions are such that the in-
jected fluids will not migrate within 10,000 years:
vertically upward out of the injection zone; or
laterally within the injection zone to a point of dis-
charge or interface with an Underground Source
of Drinking Water (USDW) as defined in 40 CFR
Part 146.
2) Before the injected fluids migrate out of the injec-
tion zone or to a point of discharge or interface
with USDW, the fluid will no longer be hazardous
because of attenuation, transformation, or immo-
bilization of hazardous constituents within the
injection zone by hydrolysis, chemical interac-
tions, or other means.
The state of the art of fluid-transport modeling is con-
siderably more advanced than that of geochemical-fate
and transport modeling. Consequently, geochemical-fate
modeling is most likely to be used if a fluid-flow no-
migration standard cannot be met. Geochemical-fate/
transport modeling of deep-well-injected hazardous
wastes is in very early stages of development, and its
use in meeting current EPA Underground Injection
Control regulations is unbroken ground. However,
where the no-migration standard must be considered,
this reference guide can help determine whether
geochemical-fate/transport modeling of a specific waste
is even feasible, and what approaches might be taken.
Organization
This reference guide follows the format of its com-
panion volume (Assessing the Geochemical Fate of
Hazardous Wastes: Summaries of Recent Research).
The contents and organization of each chapter are:
Chapter One (Overview of Deep-Well Injection of
Hazardous Wastes in the United States) discusses
the identification and properties of significant wastes
(Section 1.1). The sources, amounts, and composi-
tions of deep-well-injected wastes are summarized in
Section 1.2, and the geographic distribution in in-
dustrial injection wells is covered in Section 1.3. The
chapter concludes with a discussion of the design
and construction of such wells (Section 1.4).
Chapter Two (Processes Affecting the Geochemi-
cal Fate of Deep-Well-Injected Wastes) begins
with an overview of influences on the geochemical
fate of injected wastes (Section 2.1). This section
discusses key characteristics of chemical systems,
fate-influencing processes (partition, transformation,
and transport), and interactions between hazardous
waste and deep-well reservoirs. Subsequent sec-
tions examine partition (Section 2.2), transformation
(Section 2.3), and transport processes (Section 2.4)
in more detail. Chapter Two concludes with a discus-
sion of the interactions among partition, transforma-
tion, and transport processes in the deep-well
environment (Section 2.5).
Chapter Three (Major Environmental Factors Af-
fecting Deep-Well-Injection Geochemical Proces-
ses) examines the environmental factors that
determine what types of processes may occur and their
outcomes. Section 3.1 discusses specific environmen-
tal factors (pH, redox potential, salinity, reservoir
matrix, temperature, and pressure). Section 3.2
reviews brines and the major types of rocks in injec-
tion zones (carbonates and sandstones) and confin-
ing beds. Section 3.3 assesses the implications of
these environmental factors for well plugging, con-
fining formation failure, and well blowout. Finally,
xv
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Section 3.4 evaluates the influence of deep-well
conditions on biodegradation.
Chapter Four (Geochemical Characteristics of Haz-
ardous Wastes) begins with a general discussion of
inorganic and organic hazardous wastes (Section 4.1)
followed by more detailed information on inorganic
wastes (Section 4.2) and organic wastes (Section 4.3).
Section 4.4 contains suggestions on how to locate ad-
ditional data on a substance of interest.
Chapter Five (Methods and Models for Predicting the
Geochemical Fate of Deep-WelMnjected Wastes)
covers basic approaches (Section 5.1) and specific
methods and models (Section 5.2). The latter section
discusses computer codes for both aqueous and
solution geochemistry (Section 5.2.1), adsorption
(Section 5.2.2), biodegradation (Section 5.2.3),
hydrolysis (Section 5.2.4), and chemical transport
(Section 5.2.5).
Chapter Six (Field Sampling and Laboratory Proce-
dures and Protocols) summarizes specific laboratory
procedures for geochemical-fate assessment. This
chapter includes a discussion of how to select sampling
methods and laboratory procedures (Section 6.1.2);
methods for characterizing the waste to be injected and
the injection reservoir (Section 6.2); types of waste-
reservoir interaction tests (Section 6.3); procedures for
measuring adsorption isotherms, hydrolysis rate
constants, and biodegradation (Section 6.4); and a
brief discussion of quality-control/assurance pro-
cedures (Section 6.5). The chapter concludes
with an annotated bibliography of laboratory and
field geochemical studies of deep-well waste
injection and references related to laboratory
procedures and protocols (Section 6.6).
Chapter Seven (Case Studies of Deep-Well
Injection of Industrial Wastes) discusses the
methods for field investigation of geochemical fate
(Section 7.1) and then describes six cases where
geochemical interactions of injected wastes in a
variety of deep-well environments have been
studied (Sections 7.2 to 7.7). Each case study
contains information on (1) the injection facility,
including the waste characteristics and history
of injection activities, (2) lithology and chemistry of
the injection zone and confining layer, and
(3) geochemical processes observed or inferred to
occur in the injection zone.
How to Use This Reference Guide
Because the study of the geochemical fate of wastes
in the deep-well environment involves a range of
scientific disciplines, this guide was written so that an
expert in a field can quickly find information and ref-
erences, but that others needing more of an over-
view of processes, chemistry, modeling techniques,
laboratory procedures or other such information can
readily gain a deeper knowledge of the subject. Sug-
gestions for using this guide follow.
Information on a specific inorganic hazardous
waste. For an inorganic waste, such as lead or
chromium, turn to Section 4.2, which identifies
relevant processes. You may also want to read the
discussion of relevant processes in Chapter Two.
Information on organic hazardous wastes. For an or-
ganic waste, such as phenol, turn to the appropriate sub-
section in Section 4.3 (Monocydic Aromatics, Section
4.3.3, in the case of phenol). To determine to which or-
ganic group the substance belongs, check Appendix A.
The tables on the characteristics of each group of com-
pounds (Tables 4-5 through 4-11) list other tables in the
handbook containing information on individual com-
pounds. Appendix B contains a list of hazardous organic
compounds that have been studied in ground-water con-
taminant transport studies. Literature citations giving par-
tition coefficients or retardation factors and ground-water
biodegradation studies for the particular compound are
listed. Again, you may want to read the discussion of
relevant processes in Chapter Two, and check the index
in Appendix A for other references to the substance.
Section 4.4 lists data bases that can be used to perform
literature searches for a specific substance.
Information on specific reservoir conditions. Section
3.2 discusses brine characteristics and reservoir-confining
rocks. This section also has a reference index (Table 3-9)
that includes citations to literature on the geology in areas
of the United States where injection is practiced or the
feasibility of injection has been assessed.
Information on a specific process, environmental
factor, or fate-prediction method or model. Turn to
the appropriate sections in Chapters Two (Processes),
Three (Environmental Factors), and/or Four (Methods
and Models).
Information on laboratory procedures. Turn to
Chapter Six. The bibliography in this chapter gives
detailed information about all the citations.
XVI
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CHAPTER ONE
OVERVIEW OF DEEP-WELL INJECTION OF HAZARDOUS WASTES
IN THE UNITED STATES
This chapter discusses the characteristics of hazard-
ous wastes typically injected into Class I injection
wells. It includes:
• The properties that define a waste as hazardous
(Section 1.1)
• The sources, amounts, and composition of existing
deep-well-injected hazardous wastes (Section 1.2)
• Trends and distribution of industrial and hazardous
waste injection (Section 1.3)
• The design and construction of deep-injection
wells (Section 1.4)
1.1 Identifying Hazardous Wastes
Wastes are defined as hazardous for purposes of
regulatory control in 40 CFR Part 261. In this regula-
tion, wastes are classified as hazardous either by
being listed in tables within the regulation or by
meeting certain specified characteristics. Thus under
40 CFR Part 261 hazardous wastes are known either
as listed or characteristic wastes. Some listed
wastestreams, such as spent halogenated solvents
(listed in 40 CFR 261.31), come from many in-
dustries and processes. Other listed wastestreams,
such as API separator sludges from the petroleum-
refining industry (listed in 40 CFR 261.32), come
from one particular industry and one process. A
characteristic waste is not listed, but is classified as
hazardous because it exhibits one or more of the fol-
lowing characteristics:
• Toxicity to living organisms
• Reactivity
• Corrosivity
• Ignitability
Listed wastes also exhibit one or more of these char-
acteristics.* The significance of each of the charac-
teristics listed above is discussed below and is
summarized in Table 1-1. Deep-well-injected wastes
commonly contain several components that classify
the waste as hazardous, along with other nonhazardous
components.
1.1.1 Toxicity
A waste is toxic under 40 CFR Part 261 if the extract
from a representative sample of the waste exceeds
specified limits for eight elements and four pesticides (ar-
senic, barium, cadmium, chromium, lead, mercury,
selenium, silver, endrin, methoxychlor, toxaphene, 2,4-D
and 2,4,5-TP Silvex—see Table 1, 40 CFR 261.24)
using extraction procedure (EP) toxicity test methods.
Note that this narrow definition of toxicity relates to
whether a waste is defined as hazardous for regulatory
purposes; in the context of this reference guide, toxicity
has a broader meaning because most deep-well-injected
wastes have properties that can be toxic to living or-
ganisms.
1.1.2 Reactivity
Reactivity describes a waste's tendency to interact
chemically with other substances. Many wastes are
reactive, but it is the degree of reactivity that defines
a waste as hazardous. Hazardous reactive wastes
are those which are normally unstable and readily
undergo violent change without detonating, react
violently with water, form potentially explosive mix-
tures with water, generate toxic gases or fumes
when combined with water, contain sulfide o,'
cyanide and are exposed to extreme pH conditions,
or are explosive. Because deep-well-injected waste-
streams are usually dilute (typically less than 1°o
*Note: Radioactive wastes are not covered by 40 CFR
Part 261. They involve environmental and regulatory is-
sues that are beyond the scope of this reference guide.
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waste in water) hazardous reactivity is not a significant
consideration in deep-well injection, although individual
compounds may exhibit this property at higher con-
centrations than those which exist in the wastestream.
Nonhazardous reactivity is, however, an important
property in deep-well injection, since when a reactive
waste is injected, precipitation reactions that can lead
to well plugging may occur (see Section 3.3.1).
1.1.3Corrosivity
Corrosive wastes are defined as those wastes with a
pH < 2 > 12.5 (i.e., the waste is very acidic or very
basic). Beyond its importance in defining a waste as
hazardous , the corrosivity of wastes is also a proper-
ty of concern to deep-well injection systems and opera-
tions. Corrosive wastes may damage the injection
system, typically by electrochemical or microbiological
means. Corrosion of injection well pumps, tubing, and
other equipment can lead to hazardous waste leaking
into strata not intended for injection. For information on
various types of electrochemical corrosion relevant to
the injection-well system, the reader is referred to
Warner and Lehr (1977). Other recommended sources
include Langelier (1936), Ryznar (1944), Larson and
Buswell (1942), and Stiff and Davis (1952). These
sources discuss saturation and stability indexes for
predicting the potential for corrosion or scaling (ac-
cumulation of carbonate and sulfate precipitates) in
injection wells. The Stiff and Davis index is recom-
mended by Warner and Lehr (1977) as most applicable
to deep-well injection of hazardous wastes, because it
is intended for use with highly saline ground waters.
Additionally, Ostroff (1965) provides examples of how
to use the index, Watkins (1954) describes procedures
that test for corrosion, and Davis (1967) thoroughly dis-
cusses microbiological corrosion of metals.
1.1.4 Ignitability
As noted, deep-well-injected wastes are relatively
dilute. Therefore, Ignitability is not a significant con-
sideration in deep-well injection, although in a con-
centrated form, individual compounds may exhibit
this property. Ignitability has no further implications
for the fate of deep-well-injected waste.
1.2 Sources, Amounts, and Composition
of Deep-Well-Injected Wastes
The sources, amounts, and composition of injected
hazardous wastes are a matter of record, since the
Resource Conservation and Recovery Act (RCRA)
requires hazardous waste to be manifested (i.e., a
record noting the generator of waste, its composition
or characteristics, and its volume must follow the
Table 1-1 Typical Physical/Chemical Properties
of Hazardous Components in
Deep-Wei l-lnjected Wastes
Characteristic
Comment
Hazardous Characteristics
Toxicity
Reactivity
Corrosivity
Ignitability
Has toxic properties that
result in classification as a
hazardous waste, but
specific properties may vary
greatly.
Reactivity usually reduced
by dilution; actual
concentration may affect
toxicity and mobility.
May be a significant
consideration in well design
and geochemical fate.
Not a significant
consideration under injection
conditions.
Physical/Chemical Properties
Normal physical state
Molecular weight
Density/Specific
gravity
Solubility
Boiling point
Melting point
Vapor
pressure/Density
Flash point/
Autoignition point
Liquids or dissolved solids.
May affect structure-activity
relationships
(see Section 4.4.4).
Must be miscible in water.
Must be soluble or miscible
in water.
Greater'than ambient
temperatures.
Less than ambient
temperatures.
Water soluble volatile
compounds may be
involved, but vapor pressure
and vapor density are not
significant considerations in
deep-well injection.
Greater than ambient
temperatures.
-------
waste load from its source to its ultimate disposal
site). The sources and amounts of injected hazard-
ous waste can be determined, therefore, based on
these records. Table 1-2 shows the estimated
volume of deep-well-injected wastes by industrial
category for 1983, the most recent year for which
data summaries are available. More than 11 billion
gallons of hazardous waste were injected in 1983.
Organic chemicals (51%) and petroleum-refining and
petrochemical products (25%) accounted for three-
quarters of the volume of injected wastes that year.
The remaining 24% was divided among six other in-
dustrial categories: miscellaneous chemical products,
agricultural chemical products, inorganic chemical
products, commercial disposal, metals and minerals,
and aerospace and related industry.
Although the general composition of each shipment
of wastes to an injection well may be known, a num-
ber of factors makes it difficult to characterize fully
the overall composition of industrial wastewaters at
any one well. These factors include (1) variations in
flow, in concentrations, and in the nature of organic
constituents over time; (2) biological activity that may
transform constituents overtime; and (3) physical in-
homogeneity (soluble and insoluble compounds)
(Hunter, 1971). Further, the exact composition of the
Table 1-2 Estimated Volume of
Deep-Well-Injected Wastes by
Industrial Category, 1983
Industrial Volume
Category (million gal/yr)
Organic chemical
Petroleum refining and
petrochemical products
Miscellaneous chemical
products
Agricultural chemical products
Inorganic chemical products
Commercial disposal
Metals and minerals
Aerospace and related
industry
Total3
5,868
2,888
687
525
254
475
672
169
1 1 ,539
Percent
of Total
50.9
25.0
6.0
4.6
2.2
4.1
5.8
1.5
100.0
aTotal may not add due to rounding.
Source: U.S. EPA (1985).
shipment may not be known because of chemical
complexity (Hunter, 1971). An example of the com-
plexity of organic wastes is illustrated in Roy et al.
(1989), which presents an analysis of an alkaline
pesticide-manufacturing waste. This waste contained
more than 50 organic compounds, two-fifths of which
could not be precisely identified.
Although no systematic data base exists on the
exact composition of deep-well-injected wastes in a
survey of 209 operating waste-injection wells, Reeder
et al. (1977) found that 53% injected one or more
chemicals identified in that study as hazardous. The
U.S. EPA gathered data for 108 wells (55% of total
active wells) that were operated in 1983. Table 1-3
summarizes the total quantity of undiluted waste in
six major categories, provides a breakdown of
average concentrations of constituents for which
data were available, and indicates the number of
wells involved. A little more than half the undiluted
waste volume was composed of nonhazardous inor-
ganics (52%). Acids were the most important con-
stituent by volume (20%), followed by organics
(17%). Heavy metals and other hazardous inor-
ganics made up less than 1% of the total volume in
the 108 wells. About a third of the wells injected
acidic wastes and about two-thirds injected organic
wastes. Although the percentage of heavy metals by
volume was low, almost one-fifth of the wells injected
wastes containing heavy metals.
An injected wastestream typically is composed of the
waste material and a large volume of water. Because
the data in Table 1-3 include only 55% of the injec-
tion wells that were active in 1983, it is not possible
to estimate precisely the percentage of waste to the
total volume of injected fluid shown in Table 1-2.
However, if the same total proportions apply to all
wells, wastes made up of 3.6% of the total volume of
injected fluid (36,000 mg/L). This percentage agrees
well with an independent estimate for a typical injec-
tion ratio of 96% water and 4% waste (Strycker and
Collins, 1987).
Table 1-3 also shows that the average concentration
of all the acidic wastes exceeded 40,000 mg/L. Con-
centrations of metals ranged from 1.4 mg/L (chromium)
to 5,500 mg/L (unspecified metals, probably containing
multiple species). Five of the 18 organic constituents
exceeded 10,000 mg/L (total organic carbon, organic
acids, formaldehyde, chlorinated organics, and formic
acid); four exceeded 1,000 mg/L (oil, isopropyl alcohol,
urea nitrogen, and organic peroxides).
-------
Table 1-3 Waste Characteristics of
Waste Type/
Components
Acids
Hydrochloric acid
Sulfuric acid
Nitric acid
Formic acid
Acid, unspecified
Heavy Metals
Chromium
Nickel
Metals, unspecified
Metal hydroxides
Hazardous Inorganics
Selenium
Cyanide
Organics
Total organic carbon (TOC)
Phenol
Oil
Organic acids
Organic cyanide
Isopropyl alcohol
Formaldehyde
Acetophenone
Urea "N"
Chlorinated organics
Formic acid
Organic peroxides
Pentachlorophenol
Acetone
Nitrite
Methacrylonitrile
Ethylene chloride
Carbon tetrachloride
Nonhazardous Inorganics
Other
Total
108 Hazardous Waste Wells Active in 1983 in the
Average
Gallons" Concentration (mg/L)
44,1 40,900 (20.3)b
78,573
43,000
75,000
75,000
44,900
1,517,600(0.7)
1.4
600
5,500
1,000
89,600 (<0.1)
0.3
391
39,674,500(17.4)
11,413
805
3,062
10,000
400
1,775
15,000
650
1,250
35,000
75,000
4,950
7.6
650
700
22
264
970
118,679,700(52.0) —
22,964,600 (9.9) —
228,02 1,800C
United States
No. of Wells
35 (32.4)b
15
6
W
2
2
12
19(17.6)
1 1
1 1
2
1
4 (3.7)
o
C-
2
71 (65.7)
24
22
6
3
w
3
3
2
2
2
2
2
2
2
2
1
•(
•)
1
50 (46.3)
33 (30.5)
108
^Gallons of nonaqueous wastes before dilution and injection.
Number in parentheses is the percentage of total.
Excludes overlaps between organics and acids.
Source: U.S. EPA (1985).
-------
1.3 Geographic Distribution of
Hazardous Waste Injection Wells
The use of wells for disposal of industrial wastes
dates back to the 1930s, but this method was not
used extensively until the 1960s, when it was imple-
mented primarily in response to more stringent water
pollution control regulations (Warner and Orcutt,
1973). Table 1-4 shows the trend in the number of
industrial waste injection wells from 1967 to 1968.
Because of slight differences in definitions, precise
comparisons cannot be made for the 4 years for
which systematic data are available. The 1967 and
1973 data represent all industrial-waste injection and
may include wells that would not now be considered
Class I wells. The 1984 data, based on a survey by
U.S. EPA, include all active Class I hazardous waste
injection wells (H); hazardous waste injection wells
that have been permitted but not yet drilled, that are
under construction and that are completed but not
yet active or that have a permit pending (HP); and
hazardous waste wells that have been temporarily or
permanently abandoned (A). The 1986 data, from a
survey by the Illinois State Geological Survey
(ISGS), include hazardous waste wells (H), proposed
hazardous waste wells (HP), and Class I nonhazar-
dous waste wells (NH). (The proposed categories for
the EPA and ISGS surveys may not follow exactly
the same criteria, and the ISGS [H] category may in-
clude some or all of the abandoned wells in the EPA
survey).
Even though the totals in Table 1-4 may not be
directly comparable, the number of industrial-waste
injection wells more than doubled between 1967 and
1986. The change from 1967 to 1986 is particularly
noteworthy. The EPA survey noted that there had
been no significant increase in new injection-well
construction since 1980 (U.S. EPA, 1985); the ISGS
data would appear to indicate a dramatic increase. If
the 263-well total is asumed to include all the 41
abandoned wells in the EPA survey, then active
wells would total 222, an increase of 27 wells in 2
years, compared with an increase of 7 wells from
1982 to 1984 reported in U.S. EPA (1985). If
proposed wells in each survey are added, the net in-
crease becomes 37 wells in 2 years.
The state totals in Table 1-4 show some interesting
patterns. Class I injection wells are concentrated in
two states, Texas (112 wells) and Louisiana (70
wells), which have a total of 69% of all wells in the
1986 (H) category. The growth from 1984 to 1986
has been concentrated in Texas, with a 38% in-
crease, from 81 (H+HP+A) to 112 (H) wells. The only
other states to show a significant increase from 1984
to 1986 in the H+HP categories are Indiana (13
proposed wells) and California (7 proposed wells).
Nine states have had industrial-waste injection wells
in the past but did not have any permitted Class I
wells in 1986 (Alabama, Colorado, Iowa, Mississippi,
Nevada, North Carolina, Pennsylvania, Tennessee,
and Wyoming). One state (Washington) had a Class I
well in 1986, but no record of industrial wastewater
injection before that year. Note that the total of active
and proposed injection wells in 1973 was 278, more
than the 252 total in the EPA 1984 survey. The
states with the largest number of wells in the 1973
survey that may have been planned but not con-
structed appear to have been Kansas (30) and
Michigan (32).
Figure 1-1 shows the number of Class I wells in the
1986 survey by state, divided into EPA regions, and
also indicates the regulatory status of such wells in
each state as of 1989. A comparison of this map with
Figure 3-1 in Chapter Three shows the heavy con-
centration of hazardous waste injection wells in three
geologic basins: Gulf Coast, Illinois Basin, and the
Michigan Basin.
1.4 Design and Construction of
Deep-injection Wells
The following description of the design and construc-
tion of deep-injection wells is adapted from Donaldson
(1964), Donaldson et al. (1974), and U.S. EPA (1985).
1.4.1 Surface Equipment Used in Waste Disposal
Figure 1-2 shows the surface equipment used in a
typical subsurface waste-disposal system. Detailed
discussion of surface treatment methods can be
found in Warner and Lehr (1977). The individual ele-
ments are:
• A sump tank or an open 30,000- to 50,000-gallon
steel tank is commonly used to collect and mix
waste streams. An oil layer or, in a closed tank,
an inert gas blanket is often used to prevent air
contact with the waste. Alternatively, large,
shallow, open ponds may provide sufficient
detention time to permit sedimentation of
particulate matter. Such ponds often are equipped
with cascade, spray, or forced-draft aerators to
oxidize iron and manganese salts to insoluble
forms that precipitate in the aeration ponds.
• An oil separator is used when the waste
contains oil because oil tends to plug the disposal
formation. The waste is passed through a settling
-------
Table 1-4 Historical Trends in the Distribution of Industrial-Waste Injection Wells
Number of Wells
State
1984°
1967 1973*
H
HP
H
1986°
HP
NH
Alabama
Alaska
Arkansas
California
Colorado
Florida
Illinois
Indiana
Iowa
Kansas
Kentucky
Louisiana
Michigan
Mississippi
Nevada
New Mexico
New York
North Carolina
Ohio
Oklahoma
Pennsylvania
Tennessee
Texas
Washington
West Virginia
Wyoming
Total
—
—
4
1
2
3
9
1
2
—
24
21
—
—
1
—
—
1
1
5
1
32
—
2
—
110
5
—
1
5
2
6
7
13
1
30
3
45
32
1
1
1
4
4
9
11
9
—
74
—
7
1
170
2
1
4
2
—
4
6
8
—
5
2
60
11
1
—
—
—
—
14
6
—
—
69
—
—
—
195
-j
1 —
— 1
— —
2 —
— —
— —
— 5
— —
— 2
— —
6 5
— 11
— —
— —
— —
— —
— 4
— 1
1 1
— 3
— —
5 7
— —
— —
1 —
16 41
1
7
3
—
4
6
9
—
5
2
70
15
—
—
1
6
—
13
7
—
112
1
1
—
263
— 2
3 —
1 —
7 —
— —
— 51
— 3
13 —
1 —
— 51
— —
— 10
1 —
— 6
— —
— —
— —
— —
— 2
— 8
— —
— 24
— —
— —
— 8
26 165
aState totals include active and proposed wells and total 278; the number of active injection wells was 170 and is shown in
the total to facilitate comparison with other years.
Class I wells, H = hazardous, HP = proposed hazardous, A = abandoned or inactive, NH = nonhazardous.
Sources: Warner (1968); Warner and Orcutt (1973); U.S. EPA (1985); Brower et al. (1989).
-------
Figure 1-1 Regulatory Status of Class I Wells in the United States (Adapted from Brower et al., 1989).
Well data from 1986 survey by Brower et al. 1989; regulatory status updated with data
from U.S. EPA Office of Drinking Water/UIC Branch.
r
Primacy granted
Primacy under consideration/pending
Welts under federal regulation
No wells or primacy
6H Number of hazardous wells
1NH Number of nonhazardous wells
IP Number of proposed wells
8UC Number of wells under construction
0 No class I wells
BAN State ban on class I wells
1 US EPA region
BAN
•Primacy imminent
tank equipped with internal baffles to separate
the oil from the waste.
A clarifier removes such paniculate matter as
polymeric floes, dirt, oil, and grease. It is often a
tank or a pond in which detention time is long
enough to allow suspended particles to settle
gradually. The process also may be accelerated by
adding a flocculating agent such as aluminum
sulfate, ferric sulfate, or sodium aluminate. Tank
clarifiers are often equipped with a mechanical
stirrer, sludge rake, and surface skimmer that
continuously remove sludge and oil.
A filter is used in some cases when coagulation
and sedimentation do not completely separate
solids from the liquid waste in areas where sand
and sandstone formations are susceptible to
plugging. Filters with a series of metal screens
coated with diatomaceous earth or cartridge
filters typically are used. Where limestone
formations with high solution porosity are used
for injection, filtration is usually not required.
A chemical treater is used to inject a bactericide
if microorganisms could cause fouling of injection
equipment and plugging of the injection reservoir.
An unlined steel clear-waste tank typically is used
to hold clarified waste before injection. The tank is
equipped with a float switch designed to start and
stop the injection pump at predetermined levels.
An injection pump is used to force the waste into
the injection zone, although in very porous
formations, such as cavernous limestone, the
hydrostatic pressure of the waste column in the well
is sufficient. The type of pump is determinerj
primarily by well-head pressures required, the
volume of liquid to be injected, and the corrosiveness
of the waste. Single-stage centrifugal pumps are
-------
Figure 1-2 Typical Above-Ground Components of a Subsurface Waste Disposal System
(Donaldson, 1964).
• Oil separator
Chemical treater
Clear-watt* tank
Injection pump
To dispo»oI well
Sump'
used in systems that require wellhead pressures
up to about 150 psi, and multiplex piston pumps are
used to achieve higher injection pressures.
1.4.2 Injection-Well Construction
Most injection wells are drilled using the rotary
method, although depending on the availability of
equipment and other site-specific factors, reverse-
rotary or cable-tool drilling may be used. The con-
struction of an injection well incorporates several
important elements: (1) bottom-hole and injection-interval
completion, (2) casing and tubing, (3) packing and
cementing, (4) corrosion control, and (5) mechanical-
integrity testing. A detailed discussion of the technical
aspects of industrial-waste injection-well construction
can be found in Warner and Lehr (1977). U.S. EPA
(1985) also presents a survey of well construction
methods and materials used for 229 hazardous waste
injection wells.
Two types of injection well completions are used
with hazardous waste injection wells:
• Open hole completion typically is used in competent
formations such as limestone, dolomite, and con-
solidated sandstone that will stand unsupported in a
borehole. In 1985, 27% of Class I wells were of this
type, with most located in the Illinois Basin.
• Gravel pack and perforated completions are used
where unconsolidated sands in the injection zone
must be supported. In gravel-pack completions the
cavity in the injection zone is filled with gravel or,
more typically, a screen or liner is placed in the
injection-zone cavity before the cavity is filled with
gravel. In perforated completions, the casing and
cement extend into the injection zone and are then
perforated in the most permeable sections. In 1985,
53% of Class I wells were perforated and 17%
were screened (U.S. EPA, 1985).
Casing and tubing are used to prevent the hole from
caving in and to prevent aquifer contamination by con-
fining wastes within the well until they reach the injec-
tion zone. Lengths of casing of the same diameter are
connected together to form casing strings. Usually two- or
three-casing strings are used. The outer casing seals the
near-surface portion of the well (preferably to below the
point where aquifers containing less than 10,000 mg/L
total dissolved solids, potential underground sources of
drinking water, are located). The inner casing extends
to the injection zone. Tubing is placed inside the inner
casing to serve as the conduit for injected wastes, and
the space between the tubing and casing is usually
filled with kerosene or diesel oil after packing and
cementing are completed.
-------
Packers are used at or near the end of the injection
tubing to plug the space, called the annulus, be-
tween the injection tubing and the inner casing.
Cement is applied to the space between the outer
walls of the casing and the borehole or other casing.
Portland cement is used most commonly for this pur-
pose, although when acidic wastes are injected, spe-
cial acid-resistant cements are sometimes used in
the portion of the well that passes through the confin-
ing layers.
Corrosion control can be handled several ways:
(1) by using corrosion-resistant material in construct-
ing the well, (2) by treating the waste stream through
neutralization or other measures, and (3) by cathodic
protection.
Mechanical integrity testing is required by EPA
regulations (40 CFR 146.08[b] and [c]) to ensure
that an injection well has been constructed or is
operating without (1) significant leakage from the
casing, tubing, or packer or (2) upward movement of
fluid through vertical channels adjacent to the well
bore. Table 1-5 lists types of mechanical integrity
tests and situations in which they might be used. A
detailed discussion of mechanical integrity can be
found in U.S. EPA (1989).
References*
Brower, R. D., et al. 1989. Evaluation of Under-
ground Injection of Industrial Waste in Illinois,
Final Report. Illinois Scientific Surveys Joint Report
2. Illinois State Geological Survey, Champaign,
Illinois.
Davis, J. B. 1967. Petroleum Microbiology. Elsevier
Publishing Co., New York.
Donaldson, E. C. 1964. Subsurface Disposal of In-
dustrial Wastes in the United States. U.S. Bureau of
Mines Information Circular 8212.
*References with more than six authors are cited
with "et al."
Table 1-5 Applicability of Tests That May Be Used for Mechanical Integrity Verification
Test
Cause of Injection Well Failure
Leaks in Casing
Tubing or Packer
Presence Location
Fluid Movement
Behind Casing
Presence Location
Types
of Casing
Metal
aCan be "yes," if test staged.
bl_og response may be somewhat dampened—test may not be adequate.
°May be used with approval of EPA administrator.
dOnly if access by tracer can be gained through the casing or beneath casing shoe.
eMay indicate potential failure site by showing corrosion spots and holes.
PVC
Pressure test
Monitor annulus pressure
Temperature log
Noise log
Radioactive tracer log0
Cement bond logc
Caliper logc
Casing condition logc
yes
yes
yes
yes
yes
noe
noe
yese
noa
no
yes
yes
yes
noe
no8
yes8
no
no
yes
yes
yesd
yese
noe
noe
no
no
yes
yes
yesd
yes8
noe
no8
yes
yes
yes
yes
yes
yes
yes
yes
yes
yes
yes"
yesb
yes
yesb
yes
no
Source: U.S. EPA (1985).
-------
Donaldson, E. C., R. D. Thomas, and K. H.
Johnston. 1974. Subsurface Waste Injections in the
United States: Fifteen Case Histories. U.S. Bureau of
Mines Information Circular 8636.
Hunter, J. V. 1971. Origin of Organics from Artificial
Contamination. In Organic Compounds in Aquatic
Environments, S. D. Faust and J. V. Hunter, eds.
Marcel Dekker, Inc., New York, pp. 51-94.
Langelier, W. F. 1936. The Analytical Control of Anti-
Corrosion Water Treatment. J. Am. Water Works
Assoc. 28:1500-1521.
Larson, T. E., and A. M. Buswell. 1942. Calcium Car-
bonation Saturation Index and Alkalinity Interpreta-
tions. J. Am. Water Works Assoc. 34:1667-1684.
Ostroff, A. G. 1965. Introduction to Oil Field Water
Technology. Prentice-Hall, Englewood Cliffs, New
Jersey.
Reeder, L. R., et al. 1977. Review and Assessment
of Deep-Well Injection of Hazardous Wastes. EPA
600/2-77-029a-d, NTIS PB 269 001-004.
Roy, W. R., S. C. Mravik, I. G. Krapac, D. R. Dickerson,
and R. A. Griffin. 1989. Geochemical Interactbns of
Hazardous Wastes with Geological Formations in
Deep-Well Systems. Environmental Geology Notes
130. Illinois State Geological Survey, Champaign,
Illinois. [An earlier version of this report by the same
title was published in 1988 by the Hazardous Waste
Research and Information Center, Savoy, Illinois].
Ryzner, J. W. 1944. A New Index for Determining
Amount of Calcium Carbonate Scale Formed by
Water. J. Am. Water Works Assoc. 36:472-486.
Stiff, H. A., and L. E. Davis. 1952. A Method for
Predicting the Tendency of Oil Field Waters To
Deposit Calcium Carbonate. Am. Inst. Mining Metall.
Engineers Trans, Petroleum Div. 195:213-216.
Strycker, A., and A. G. Collins. 1987. State-of-the-Art
Report: Injection of Hazardous Wastes into Deep
Wells. EPA/600/8-87/013. NTIS PB87-170551.
U.S. Environmental Protection Agency. 1985. Report
to Congress on Injection of Hazardous Wastes. EPA
570/9-85-003, NTIS PB86-203056.
U.S. Environmental Protection Agency. 1989. Injection
Well Mechanical Integrity Testing. EPA 625/9-89/007.
Warner, D. L. 1968. Subsurface Disposal of Liquid
Industrial Wastes by Deep-Well Injection. In Subsur-
face Disposal in Geologic Basins—A Study of Reser-
voir Strata, J. E. Galley, ed. Am. Assn. Petr. Geol.
Mem. 10, pp. 11-20.
Warner, D. L., and J. H. Lehr. 1977. An Introduction
to the Technology of Subsurface Waste water Injec-
tion. EPA 600/2-77-240, NTIS PB279 207.
Warner, D. L., and D. H. Orcutt. 1973. Industrial
Wastewater-lnjection Wells in United States—Status
of Use and Regulation, 1973. In Symposium on Un-
derground Waste Management and Artificial Recharge,
J. Braunstein, ed. Pub. No. 110, Int. Assn. of Hydrologi-
cal Sciences, pp. 687-697.
Watkins, J. W. 1954. Analytical Methods of Testing
Waters to be Injected into Subsurface Oil-Productive
Strata. U.S. Bureau of Mines Report of Investigations
5031.
10
-------
CHAPTER TWO
PROCESSES AFFECTING THE GEOCHEMICAL FATE OF DEEP-WELL-INJECTED
WASTES
This chapter examines the major processes that af-
fect the fate of deep-well-injected hazardous wastes.
The focus is on processes that (1) are known to
occur in the deep-well environment or (2) have not
been directly observed but are theoretically possible.
Section 2.1 provides an overview of the types and
characteristics of chemical processes that affect in-
jected wastes and the chemical interactions that can
occur between a reservoir rock and fluids. Sub-
sequent sections in this chapter provide more details
on specific partition (Section 2.2), transformation (Sec-
tion 2.3), and transport processes (Section 2.4), and
the combined effects of these processes on the
movement of injected wastes (Section 2.5).
2.1 Overview of Fate-influencing
Processes in Chemical Systems
The section provides a basic understanding of how
chemical systems and geochemical processes
operate. Included are:
• The key characteristics of chemical systems
(Section 2.1.1)
• The three major types of fate-influencing proces-
ses that affect wastes in the deep-well environ-
ment (Section 2.1.2)
2.1.1 Key Characteristics of Chemical Systems
A chemical system is a mixture of individual com-
ponents. Chemical systems can be described by in-
teractions that occur within the system and by the
effect these processes have on the chemical com-
position and phases of the system. Interactions that
change the chemical structure of system com-
ponents are called chemical reactions. (Other inter-
actions, such as processes that alter the solubility of
system components, change the system without al-
tering chemical structures.) Whether one reaction or
a set of reactions occurs and how quickly the reac-
tions proceed are determined by the thermodynamics
and kinetics of the system. Section 2.1.1.2 discusses
thermodynamics, kinetics, and equilibrium of chemical
systems in general; Section 2.1.1.3 discusses the key
characteristics of chemical reactions.
2.1.1.1 Phases and Speciation
A substance may exist in one of three phases—solid,
liquid or gas. The mobility of a substance in the sub-
surface is influenced by which of several form or
species it may take. Species in deep-well injection
formations, fall into six main categories:
1. "Free" tons are surrounded only by water molecules
and are very mobile in ground water. Acid-base (Sec-
tion 2.2.1) and dissolution reactions (Section 2.2.3)
create free ions.
2. Species with low solubility in water may exist in solid
form (such as AgaS, BaSO-t) or liquid form (such
as chlorinated solvents). Precipitation reactions
(Section 2.2.3) and immisicible-phase separation
(Section 2.2.4) are important processes affecting this
type of speciatbn.
3. Metal/ligand complexes (such as AI[OH]+2, Cu-
humate) and organic/ligand complexes tend to
be mobile in ground water (see Section 2.3.2).
4. Physically adsorbed species are immobile in
ground water but may be remobilized if replaced
by other species with a stronger affinity to the
solid surface (see Section 2.2.2).
5. Species held on a surface by ion exchange
(such as calcium ions on clay) are also immobile
in ground water. As with physically adsorbed
species, they may be replaced by ions with a
greater affinity to the solid surface.
6. Species may differ by oxidation state—for ex-
ample, manganese (II) and (IV); iron (II) and (III);
and chromium (111) and (VI). Oxidation state is
11
-------
influenced by the redox potential (see Section 2.3.4).
Mobility is affected because oxidation state in-
fluences precipitation-dissolution reactions (Sec-
tion 2.2.3) and also toxicity in the case of heavy
metals.
Dissolved species may be ionic or nonionic. In ionic
species, an excess or shortage of electrons in the
chemical structure creates a net positive or negative
charge. In nonionic species, all negative and posi-
tive charges cancel each other out to form a neutral
molecule. Cations are positively charged ions (Na+,
Ca+2) and anions are negatively charged (SO
-------
deep-well injection, will be discussed in detail in Sec-
tion 2.2 and Section 2.3.1
Homogeneous reactions in the deep-well environ-
ment take place in only one phase (aqueous). These
reactions generally occur uniformly throughout the
phase and are easier to study and predict than
heterogeneous reactions. Heterogeneous reactions
(for example, adsorption) tend to occur at the
interface between phases. Some reactions
(such as precipitation) may result in phase changes.
Heterogeneous reactions also tend to occur more ac-
tively at some locations in the chemical system than
at others. Bacterial decomposition of wastes is a
heterogeneous process that will be more active in
locations with conditions favorable to organisms and
less active elsewhere.
The reversibility of reactions is another important
characteristic in assessing the fate of deep-well-injected
wastes. Depending on environmental conditions, re-
versible reactions readily proceed in either or both
directions. Most acid-base reactions exemplify revers-
ible processes. In aqueous solutions, relatively minor
changes in such factors as pH or concentration can
change the direction of these reactions. Irreversible
reactions, typified by hydrolysis, have a strong tenden-
cy to go in one direction only.
Table 2-1 lists the reversible and irreversible processes
that may be significant in the deep-well environment.
The characteristics of the specific wastes (see Chapter
Four) and the environmental factors present in a well
(see Chapter Three) strongly influence which processes
will occur and whether they will be irreversible. Irre-
versible reactions are particularly important. Waste
rendered nontoxic through irreversible reactions may
be considered permanently transformed into a non-
hazardous state. Rubin (1983) provides a systematic
discussion of mathematical modeling of ground-water
chemical transport by reaction type.
2.1.2 Fate-Influencing Processes in the
Deep-Well Environment
At the simplest level, the processes that most in-
fluence geochemical fate can be divided into three
groups: partition, transformation, and transport.
Only the first two are primarily geochemical in na-
ture; these will be discussed in detail. Transport is
discussed only briefly; a thorough discussion is
beyond the scope of this reference guide.
• Partition processes affect the form or state of a
specific chemical substance at a given time or
under specific environmental conditions, but not
its chemical structure or toxicity. Thus, a sub-
stance may be in a solid form or in solution
(described by the precipitation-dissolution process),
but its toxicity remains unaltered regardless of form.
The form or state of a substance, however, influen-
ces the transformation and transport processes that
can occur. For this reason, partition processes are
important to define in a fate assessment.
• Transformation processes alter the chemical
structure of a substance. In the deep-well environ-
ment, the transformation processes that may occur
are largely determined by the conditions created by
partition processes and the prevalent environmental
factors. Transport processes do not need to be con-
sidered if transformation processes irreversibly
change a hazardous waste to a nontoxic form.
• Transport processes carry wastes through the
subsurface environment and must be considered in
a fate assessment if the interaction of partition and
transformation processes does not immobilize or
alter the hazardous waste. Waste migration can
take place either in solution or in solid form (particle
migration).
Table 2-2 presents the partition and transformation
processes known to occur in the near-surface en-
vironment along with the special factors that should
be considered when evaluating data in the context of
the deep-well environment. Geochemical processes
affecting hazardous wastes in deep-well environ-
ments have been studied much less than those oc-
curring in near-surface environments (such as soils
and shallow aquifers). Consequently, laboratory data
and field studies for a particular substance may be
available for near-surface conditions, but not for
deep-well conditions.
Most of the processes listed in Table 2-2 are dis-
cussed in Sections 2.2 and 2.3. Two significant
transformation processes that affect many hazardous
organic wastes in the near-surface environment but
do not occur in deep wells are volatilization (change
from liquid to gaseous phase) and photolysis
(decomposition in sunlight). Neither of these processes
will be explored further in this reference guide.
As Table 2-2 shows, several processes can occur in
both the near-surface and deep-well environments.
For example, neutralization of acidic or alkaline wastes
is a straightforward process, and although temperature
differences between the two environments may need
to be considered, no other factors make the deep-
well setting distinctly different. The same holds true
for oxidation-reduction (redox) processes.
13
-------
Table 2-2 Near-Surface Geochemical Processes and Their Relevance to the Deep-Well Environment
Process
Surface Data
Applicable to
Deep-Well
Environment?
Comments
Partition Processes
Acid-base
equilibria
Adsorption-
desorption
Precipitation-
dissolution
Immiscible-
phase
separation
Partly
Partly
Partly
No
Complexation
Hydrolysis
Neutralization
Oxidation-
reduction
Partly
Partly
Partly
Partly
Near-surface studies tend to investigate fresh or moderately saline water,
which creates quite different conditions for acid-base equilibria. Studies of
ocean geochemistry come closest to approximating deep-well conditions.
Mechanisms for adsorption on similar materials will be similar. Soil-
adsorption data generally do not reflect the saturated conditions of the
deep-well environment. Organic-matter content is a major factor affecting
adsorption in the near-surface; its significance in the deep-well environment
is less clear. Fate studies involving artificial recharge are probably useful,
but differences between fresh waters and deep brines may reduce relevance.
Higher temperatures, pressures, and salinity of the deep-well environment
may result in significant differences between reactions in the
two environments.
Fluids (such as gasoline) that are immiscible in water are a significant
consideration in near-surface contamination. Deep-well injection is generally
limited to waste streams that are soluble in water. Well blowout from
gaseous carbon-dioxide formation is an example of this process that is
distinct to the deep-well environment.
Transformation Processes
Volatilization
Photolysis
Biodegradation
No
No
Partly
No atmosphere.
No sunlight.
Some near-surface bacteria appear <
capable of entering and surviving in the
deep-well environment. However, in general, temperature and pressure
conditions in the deep-well environment are unfavorable for microbiota that
are adapted to near-surface conditions. Biological transformations are
primarily anaerobic.
Humic substances are very significant factors in near-surface complexation
processes, probably less so in the deep-well environment. Data on
complexation in saline waters are probably most relevant.
Basic processes will be the same. Higher salinity of deep-well environment
may affect rate constants.
Basic process is the same, but some adjustments may be required for
pressure/temperature effects.
The deep-well environment tends to be more reducing than the near-surface
environment, but equally reducing conditions occur in the near-surface.
Some adjustments may be required for pressureAemperature effects.
14
-------
The remaining processes, although they occur under
both near-surface and deep-well conditions, are less ap-
plicable to the latter. Distinct differences between the two
environments, however, can lead to significant differen-
ces in how the processes affect a specific hazardous
substance. Compared with the near-surface environ-
ment, the deep-well environment is characterized by
(1) higher temperatures, pressures, and salinity, and
(2) lower organic matter content and Eh (oxidation-
reduction potential—see Section 2.3.4). Chapter
Three discusses the significance of these environ-
mental factors.
Table 2-3 lists the partition and transformation
processes applicable in the deep-well environment
and indicates whether they significantly affect the
toxicity and/or mobility of hazardous wastes. None of
the partition processes results in detoxification
(decomposition to harmless inorganic constituents),
but all affect mobility in some way. All transformation
processes except complexation can result in
detoxification; however, because transformation
processes can create new toxic substances, the
mobility of the waste can be critical in all processes
except neutralization.
Table 2-3 also indicates whether a process is biotic
(mediated or initiated by organisms in the environ-
ment), abiotic (not involving biological mediation), or
both. Biotic processes are limited to environmental
conditions that favor growth of mediating organisms.
Abiotic processes occur under a wide range of condi-
tions. Adsorption, precipitation, complexation, and
neutralization are abiotic; all other processes in
Table 2-3 may be either.
2.2 Partition Processes
Partition processes determine how a substance is
distributed among the liquid, solid, and gas phases
and determine the chemical form, or species of a
substance (see Section 2.1.1.1).
Partitioning usually does not affect the toxic proper-
ties of the substance. Partitioning can, however, af-
fect the mobility of the waste, its compatibility with
the injection zone, or other factors that influence fate
in the deep-well environment. The major partition
processes are:
• Acid-base reactions
• Adsorption-desorption
• Precipitation-dissolution
• Immiscible-phase separation
Table 2-3 Significance of Chemical Processes in the Deep-Well Environment
Process
Detoxification
Mobility
Biotic/Abiotic
Partitioning
Acid-base equilibrium
Adsorption-desorption
Precipitation-dissolution
Immiscible-phase separation
No
No
No
No
Yes
Yes
Yes
Yes
Both
Abiotic
Abiotic
Both
Transformation
Biodegradation
Complexation
Hydrolysis
Neutralization
Oxidation-reduction
Yes
No
Yes
Yes
Yes
Yes
Yes
Yes
No
Yes
Biotic
Abiotic
Both
Abiotic
Both
15
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2.2.1 Acid-Base Reactions
Acid-base reactions affect pH (the concentration of
hydrogen ions in solution), which is a controlling factor
in the type and rate of many other chemical reactions
(see Section 3.1.1). Neutralization, a special type of
acid-base reaction that functions as a transformation
process, is discussed in Section 2.3.1.
Acids dissociate in solution yielding hydrogen ions
and anions according to the general reaction:
HA (neutral) <—> H* (cation) + A'(anion)
The ionization is reversible. The anion (acting as a
weak base) can recombine with the hydrogen ion to
re-form neutral HA. Both reactions occur continuous-
ly in solution, with the extent of ionization dependent
on the strength of the acid. Strong acids, such as
HCI, ionize completely in dilute aqueous solution.
Thus a 0.01 molar (10~2 molar) solution has a pH of
2. Weak acids, such as acetic and other organic
acids, ionize only slightly in solution and form solu-
tions with pH from 4 to 6.
In the above example, the anion (A") functions as a
base when it combines with a hydrogen ion. (By
definition, any substance that combines with
hydrogen ions is a base. Like strong acids, strong
bases ionize completely in a dilute aqueous solu-
tion.) Thus NaOH dissolves in water to form
hydroxide ions, which in turn function as a base
when they combine with hydrogen ions to form
water, as shown by the general equation:
MOH<—> M+ + OH'
Strong acids (those which ionize completely in solu-
tion) are more likely to dissolve solids because
charged particles such as hydrogen ions will interact
more strongly with solids than will neutral particles.
Weak acids do not readily donate hydrogen ions and
consequently remain mostly in the neutral form. As a
result, weak acids do not dissolve solids as readily
as strong acids.
Strong bases (those which most readily extract hydrogen
bns from solution) are also found predominantly in ionic
forms and are similarly more reactive with solids than
weak bases, which remain mostly in neutral form. The
extent to which any base will extract hydrogen ions from
solution depends on pH and the strength of the base.
Acid-base reactions occur quickly. When the pH of a
solution changes, acids and bases readily attain a
new equilibrium between neutral and ionic forms. Be-
cause toxic organics almost always exist in very low
concentrations and tend to be weak acids or weak
bases, they have little, if any, influence on the pH of
water. Acid-base equilibrium reactions involving haz-
ardous organic compounds do not affect the toxicity
of the waste and, as noted above, do not strongly in-
fluence pH. Table 2-4 identifies some acidic and
basic hazardous organic wastes. Mills et al. (1985)
describe the procedures for calculating the fraction of
a toxic organic acid or base that is in nonionic,
neutral form. Although this procedure is useful
primarily for evaluating the volatilization of organics
in near-surface conditions (because only electrically
neutral species are directly volatile), it may also be
useful when evaluating adsorption behavior in the
deep-well environment.
When weak acids and bases ionize in wastestreams,
pH is affected very little, but when strong acids and
bases ionize in wastestreams, pH is affected
dramatically. By definition, wastestreams having a
pH < 2 (highly acidic) or a pH > 12.5 (strongly basic)
are highly corrosive and are regulated as hazardous.
As discussed in Section 2.3.1, acid-base reactions
can neutralize acidic or basic hazardous waste by
raising or lowering its pH.
2.2.2 Adsorption and Desorption
Adsorption is a physicochemical process whereby
ionic and nonionic solutes become concentrated
from solution at solid-liquid interfaces. Adsorption
and desorption are caused by interactions between
and among molecules in solution and those in the
structure of solid surfaces. Adsorption is a major
mechanism affecting the mobility of heavy metals
and toxic organic substances and is thus a major
consideration when assessing transport. Because
adsorption usually is fully or partly reversible
(desorption), only rarely can it be considered a
Table 2-4 Acid-Base Characteristics of Toxic
Organics
Acidic
Basic
Phenol
2-Chlorophenol
2,4,-Dichlorophenol
2,4,6-Trichlorophenol
Pentachlorophenol
2-Nitrophenol
4-Nitrophenol
2,4-Dinitrophenol
2,4-Dimethylphenol
4,6-Dinitro-o-cresol
Benzidine
Dimethylnitrosamine
Diphenylnitrosamine
Di-n-propyl nitrosamine
Source: Adapted from Mills et al., (1985).
16
-------
detoxification process for fate-assessment purposes.
Although adsorption does not directly affect the
toxicity of a substance, the substance may be
rendered nontoxic by concurrent transformation
processes such as hydrolysis and biodegradation.
Many chemical and physical properties of both
aqueous and solid phases affect adsorption, and the
physical chemistry of the process itself is complex.
For example, adsorption of one ion may result in
desorption of another ion (known as ion exchange).
Adsorption is typically exothermic (i.e., releases
energy in the process of bonding), but can be en-
dothermic, and can be classified into two groups,
based on the energies involved: chemical adsorption
and physical adsorption. Chemical adsorption is
more significant for heavy metals, either in the form
of ion exchange or interactions involving metal com-
plexes. Physical adsorption is more significant for
hazardous organic compounds and is discussed in
Section 2.2.2.2.
In chemical adsorption (also called chemisorption),
chemical bonds are formed between the adsorbate
molecule and the adsorbent. These bonds typically
involve energies on the order of 7 kcal/mole or
greater (Roy et al., 1987). These energies distinguish
them from physical bonds, which typically involve
energies less than 7 kcal/mole. Ion exchange, ligand
exchange, protonation, and hydrogen bonds typically
fall in the category of chemical bonds (see Table 2-5).
Depending on the classification scheme used,
numerous distinct types of chemical bonds have
been identified in the laboratory under controlled
conditions. Determining bonding mechanisms in the
natural environment is much more difficult because
of the diversity and complexity of adsorption sur-
faces. Chemical adsorption bonds are described
below.
2.2.2.1 Chemical Bonding Mechanisms
Most interactions between heavy metals and/or or-
ganic species and clays involve one or more of three
types of chemical bonds (Mortland, 1985): ion ex-
change, protonation, and hydrogen bonds. Where
complex molecules are involved (see Section 2.3.2),
ligand exchange may also serve as an important
bonding mechanism (Roy et al., 1987). Table 2-5
summarizes the forces, adsorbate characteristics,
and energies of these and other less common types
of bonds, which are discussed in the following para-
graphs.
Ion exchange. Ionic bonds bind metal and organic
cations to negative electrical charges on the adsor-
bent surface. The negatively charged sites where
ionic adsorption occurs are called exchange sites,
and adsorbed cations that may be displaced by other
cations are called exchangeable tons. In the deep-well
environment, most exchange sites are filled primarily
by such cations as Na+, Ca+2, and Mg+2. Consequent-
ly, any ionic chemisorption of injected toxic wastes
results from the displacement of these cations already
adsorbed. Heavy metals have a strong tendency
towards ion exchange. Ion exchange does not occur
with nonionic toxic organics and generally is insig-
nificant even for ionizable organic species because
most organics ionize only slightly in solution.
Cations that form high-energy bonds will displace
those which form lower-energy bonds. Thus, in ion
exchange, divalent metal ions (ions with two avail-
able electrons) such as Ca+2 and Mg+2, which have
a stronger charge and thus a stronger attraction to
negatively charged sites (see Section 3.1.4.2), will
displace monovalent metal ions (ions with one avail-
able electron) such as Na+ and K+. Similarly,
monovalent ions tend to displace organic molecules
that adsorb using lower-energy physical forces.
Protonation. Protonation, which may take place
without ion exchange, occurs when an acid-base
reaction takes place with an exchangeable hydrogen
ion located at the adsorption site. A neutral organic
molecule that can act as a base may protonate (i.e.,
add a hydrogen ion to its structure) and become at-
tached to the site where the H+ ion has already been
adsorbed. Protonation may occur with hydrogen ions
that have adsorbed directly onto the mineral surface
or with those attached to hydrated metal cations that
have adsorbed onto the surface.
Hydrogen bonds. The negative pole of a polar organic
molecule may be attracted to hydrogen atoms either on
the surface of complex molecules (see Section 2.3.2) or
on molecules already adsorbed onto the mineral surface.
A water bridge may link the polar organic molecule to a
water molecule on a hydrated exchangeable metal cat-
ion (see also Section 2.3.2). Similarly, a hydrogen atom
in an adsorbed organic cation may provide a site for
hydrogen bonding. In either case the shared hydrogen
atom is more strongly bound to the adsorbed molecule,
so hydrogen bonds are weaker than protonation bonds.
Ligand exchange. A ligand bond is formed when an
ion or molecule attaches to a central ion to form a
complex ion (Hamaker and Thompson, 1972). The
ion or molecule attached to the central molecule is
called a ligand. Because complex molecules tend to
be strongly adsorbed on solid surfaces, if a molecule
17
-------
(usually a water molecule) replaces a ligand, it be-
comes part of the already-adsorbed molecule.
Other bonds. The literature describes additional
types of bonds that do not appear significant in the
deep-well environment. These bonds are described
below, with citations to detailed references.
• Covalent bonds. Covalent bonding involves the
sharing of a pair of electrons by two atoms. Covalent
bonding between silicates and organic groupings
has been created in the laboratory but the degree to
which such reactions occur in soils and sediments is
unknown. The relatively high pressures and
temperatures in deep-well environments over
geologic time may cause some covalent bonding be-
tween organic matter and silicates (Mortland, 1970).
Covalent bonds have high energies on the order of
100 kcal/mole. Thus, covalent bonding probably
would not be a significant process during short-term
interactions of organic wastes and reservoir solids,
but may be important for metals.
• Charge transfer. Charge-transfer bonds result
from the transfer of electrons across the surface of
the adsorbent or organic molecules (Hamaker and
Thompson, 1972). Hydrogen bonding, discussed
above, is a special case of this phenomenon.
• Anionic exchange or adsorption. Anionic ad-
sorption has been observed on dry clay films in the
laboratory. It is not very likely, however, that or-
ganic anions will be absorbed on clays in the
saturated conditions that exist in the deep-well en-
vironment unless they possess other properties.
For example, in the case of anionic polymers with
large molecular weights, entropy-generating forces
(see hydrophobic forces in Section 2.2.2.2) might
favor adsorption (Mortland, 1985).
• Pi bonds. Pi bonding occurs when the pi electrons
of an organic compound are donated to a metal. This
bond has been observed in the laboratory between
benzene, xylene, toluene, and chlorobenzene and
Cu(ll) saturated montmorillonite clay (Doner and
Mortland, 1969), but it does not seem likely to be sig-
nificant in the deep-well environment.
2.2.2.2 Physical Adsorption Forces
Several physical forces influence adsorption. However,
only two—van der Waals and hydrophobic (water-
avoiding) forces—are significant in the deep-well en-
vironment.
Table 2-5 summarizes key information about these
forces, which with two forces of lesser importance
are discussed below.
Van der Waals forces. These are physical forces that
operate between and among all atoms, ions, and
molecules. They are relatively weak and decrease
rapidly with distance. Van der Waals interactions are
additive and can become significant when large-
molecular-weight organic compounds are present.
They also tend to be the most significant forces affect-
ing adsorption of nonpolar organic molecules. The term
van der Waals-London is also used to describe this
kind of bonding.
Hydrophobic forces. Hydrophobic forces cause
water molecules to be displaced when organic
molecules interact with the surface of an adsorbent.
This force does not result primarily from the attrac-
tion between the adsorbing surface and the organic
molecule. Rather, it occurs because the structure of
water is less stable when mixed with nonpolar or-
ganic molecules (neutral molecules in which the
electrons are uniformly distributed on the surface).
The hydrophobic interactions of nonpolar organic
molecules with water tend to "push" the organic
molecules to nearby mineral surfaces where van der
Waals forces can cause adsorption, displacement of
water molecules, and an increase in entropy. The term
entropy generation has also been used for this type of
adsorption (Mortland, 1970; Jury, 1986). The net effect
is that the system is thermodynamically more stable
when nonpolar molecules displace water molecules at
the solid surface (Hamaker and Thompson, 1972).
Dipole-dipole interactions, lon-dipole and dipole-
dipole interactions occur between polar organic
molecules and electrically charged or polar-adsorbing
surfaces. Adsorbed metal cations have the greatest
potential for providing bonding sites in the deep-well
environment. Polar organic molecules must compete
with more abundant water molecules, which are also
polar; therefore dipole-dipole interactions are probably
not major processes in ground-water systems
(Mortland, 1985).
Magnetic forces. Ring structures containing con-
jugated double bonds create magnetic currents;
thus, magnetic forces between large molecules and
those between organic humic substances with con-
jugated double bonds may be significant (Hamaker
and Thompson, 1972). But such conditions, if occur-
ring at all in the deep-well environment, would be
rare.
18
-------
Table 2-5 Major Intermolecular Interactions Involved in Adsorption in the Deep-Well Environment
Type of
Interaction
Forces
Adsorbate
Energy
(kca I/mole)
Sources
Primarily Chemical Bonds (>7 kca I/mole)
Ion exchange Electrostatic
Ligand exchange Electrostatic
Protonation Electrostatic
Hydrogen
Electrostatic
Metal cations
Organic acid/
Cation
Complex ion
Organic bases,
metals
Polar organic
up to 50
1.9-19.3
up to 35
0.5-15
1.9-10
Hamaker and Thompson, 1972
Royetal., 1987
Mortland, 1970
Hamaker and Thompson, 1972
Royetal., 1987
Primarily Physical Adsorption Forces
Van der Waals Electrostatic
Small molecules
1-2
0.03-1.9
Hydrophobic
Dipole
Magnetic
Entropy
generation
Electrostatic
Magnetic
Large molecules 11 +
Nonpolar organic ~1
Polar organic
0.6
Hamaker and Thompson, 1972
Roy et al., 1987
Hamaker and
Thompson, 1972
ibid.
ibid.
Royetal., 1987
2.2.2.3 Reversibility of Adsorption
Adsorption is often fully reversible, as can be seen in
an adsorption-desorption cycle. First, a mineral is
exposed to a solution with a known concentration of
a compound until equilibrium is reached. Then the
same mineral is exposed to the same solution (but
without the added solute) until no further desorption is
noted. If adsorption is a fully thermodynamic process,
the amount of the compound adsorbed will equal the
amount that is desorbed; that is, all adsorbed material
is desorbed.
A number of investigators of adsorption-desorption
behavior of pesticides on soil have observed,
however, apparent irreversibility (Van Genuchten et
al., 1974). Rao and Davidson (1980) have identified
three major causes of irreversibility in laboratory ex-
periments involving adsorption:
Artifacts created by some aspect of the
laboratory method. For example, desorption ex-
periments typically involve repeated use of a
centrifuge to separate "equilibrated" solutions
from the solids, and then follow with agitation and
resuspension of solids in the solution. This proce-
dure may break down soil particles, thus increas-
ing the number of adsorption sites during the
desorption phase.
Failure to establish complete equilibrium
during adsorption. For example, slow solvent
action of the aqueous solution might unmask new
adsorption sites. Also, a pseudo-equilibrium (ap-
pearance of a steady state before true equi-
librium is attained) may result when clays or
organic matter have adsorption sites within the
particles that are reached only after slow dif-
fusion.
19
-------
• Chemical transformations of the adsorbed
substance. These transformations result in a
chemical structure that is more strongly bound to
the solid surface, or they result from microbial
degradation. Since the total amount of the sub-
stance in the system is reduced, desorption
results in concentrations that are lower than the
beginning concentration.
In spite of considerable research in this area, the
physicochemical basis for "irreversible" adsorption is
not well understood (Rao and Davidson, 1980). For
example, in flow-through adsorption experiments in-
volving phenol interacting on a Frio sandstone core
under simulated deep-well temperatures and pres-
sures, Collins and Crocker (1988) observed no
desorption when the core was flushed with brines
that did not contain phenol.
2.2.3 Precipitation and Dissolution
Precipitation is a phase-partitioning process whereby
solids separate from a solution. Dissolution involves
movement from the solid or gaseous phase to the
aqueous phase. Solids dissolve into ions, whereas
gases retain their original chemical structure when dis-
solved. The solubility of a compound (its tendency to
dissolve in water or other solutions) is the main proper-
ty affecting the precipitation-dissolution process. This
section examines the solubility characteristics of haz-
ardous wastes and the significance of precipitation and
dissolution processes in the deep-well environment.
The concentration of a compound in water is controlled
by its equilibrium solubility or solubility constant
(the maximum amount of a compound that will dis-
solve in a solution at a specified temperature and
pressure). Equilibrium solubility will change with en-
vironmental parameters such as temperature, pres-
sure, and pH; for example, the solubility of most
organic compounds triples when temperature rises
from 0°C to 30°C. Each type of waste has a specific
equilibrium solubility at a given temperature and
pressure. The solubility of toxic organic compounds
is generally much lower than that of inorganic salts.
This characteristic is particularly true of nonpolar
compounds because of their hydrophobic character
(see Section 2.2.2.2).
Precipitation usually occurs when the concentration
of a compound in solution exceeds the equilibrium
solubility, although slow reaction kinetics may result
in "supersaturated" solutions. For organic wastes in
the deep-well environment, precipitation is not
generally a significant partitioning process; in certain
circumstances, however, it may need to be con-
sidered. For example, pentachbrophenol precipitates
out of solution when the solution has a pH < 5 (Choi
and Aomine, 1974a, 1974b), and polychlorophenols
form insoluble precipitates in water high in Mg+2 and
Ca+2 ions (Davis, 1967). Also, organic anions react
with such elements as Ca , Fe+2, and Al+3 to
form slowly soluble to nearly insoluble com-
pounds.
Precipitation may be significant for heavy metals and
other inorganic constituents in injected wastes. For
example, sulfide ions have a strong affinity for metal
ions, precipitating as metal sulfides. The dissolved
constituents in injected wastes and reservoir fluids
would not be in equilibrium with the in situ brines be-
cause of the fluids' different temperature, pH, and Eh
(oxidation-reduction potential; see Section 3.6.2).
When the fluids are mixed, precipitation reactions
can lead to injection-well plugging. Section 3.3.1.3
(Well Plugging) examines specific inorganic
precipitation reactions that may cause problems
during injection.
Coprecipitation is a partitioning process whereby
toxic heavy metals precipitate from the aqueous
phase even if the equilibrium solubility has not been
exceeded. This process occurs when heavy metals
are incorporated into the structure of silicon,
aluminum, and iron oxides when these latter com-
pounds precipitate out of solution (Fisher et al.,
1974, as cited by Scrivner et al., 1986). Iron
hydroxide collects more toxic heavy metals
(chromium, nickel, arsenic, selenium, cadmium, and
thorium) during precipitation than aluminum hydroxide
(Bunshah, 1970). Coprecipitation is considered to
remove effectively trace amounts of lead and
chromium from solution in injected wastes at New
Johnsonville, Tennessee (Scrivner et al., 1986).
Coprecipitation with carbonate minerals may be an
important mechanism for cobolt, lead, zinc and cad-
mium (Forstner and Wittmann, 1979).
Dissolution of carbonates (acidic wastes), sand
(alkaline wastes), and clays (both acidic and alkaline
wastes) can neutralize deep-well-injected wastes (Scriv-
ner et al., 1986). Neutralization is discussed in Section
2.3.1. Because precipitation-dissolution reactions are
highly dependent on environmental factors such as pH
and Eh, changes in one or more factors as a result of
changes in injected-waste characteristics, or varying per-
centages of injected waste and reservoir fluids con-
centrations, may result in re-solution or reprecipitation of
earlier reaction products. This sensitivity to environmen-
tal factors increases the complexity of predicting
precipitation-dissolution reactions because different equi-
librium solubilities of a compound may exist in different
parts of the injection zone depending on the proportions
20
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of waste and reservoir fluid. Similarly, a sequence of
precipitation and dissolution reactions may take
place at a given location of the injection zone as the
concentration of injected wastes increases.
2.2.4 Immiscible-Phase Separation
An insoluble liquid or gas will separate from water,
resulting in immiscible-phase separation. The behavior
of nonaqueous-phase liquids (NAPLs) that may be
lighter (LNAPLs) or denser (DNAPLs) than water is
important in near-surface ground-water contamina-
tion studies (Palmer and Johnson, 1989). However,
aqueous-phase separation generally is not an issue
in the deep-well environment because injected haz-
ardous wastes are usually dilute. Failure to remove
immiscible oily fluids from injected wastes potentially
may cause plugging in the injection zone. Density
and viscosity differences between injected and reser-
voir fluids, however, may need to be considered in
transport modeling (see Section 2.4.4). Generally,
pressures are high enough in the deep-well environ-
ment to keep gases such as carbon dioxide,
generated as products of waste-reservoir interac-
tions, in solution. Under certain conditions of high
temperature and high waste concentrations, however,
injected hydrochloric acid can cause carbon dioxide to
separate from the liquid and produce a well blowout.
This reaction is discussed further in Section 3.3.3 (Well
Blowout) and in the case study described in Sec-
tion 7.6.
2.3 Transformation Processes
Transformation processes change the chemical struc-
ture of a compound. Because not all transformation
processes convert hazardous wastes to nonhazardous
compounds, geochemical fate assessment must con-
sider both the full range of transformation processes
that may occur and the toxicity and mobility of the
resulting products. For deep-well-injected wastes,
transformation processes and subsequent reactions
may lead to one or more of the following:
• Detoxification
• Transtoxification
• Toxification
Detoxification is an irreversible change in a sub-
stance from toxic to nontoxic form. For example,
when an organic substance breaks down into its in-
organic constituents, detoxification has taken place.
Transtoxification occurs when one toxic compound
is converted into another toxic compound. Toxifica-
tion is the conversion of a nontoxic compound to a
toxic substance. Table 2-6 lists some examples of
each. Transformation processes that may be sig-
nificant in deep-well-injection fate assessments are:
• Neutralization
• Complexation
• Hydrolysis
• Oxidation-reduction
• Catalysis
• Polymerization
• Thermal degradation
• Biodegradation
Two other processes that may transform hazardous
wastes are photolysis and volatilization, but they are
not covered here because they do not occur in the
deep-well environment (see Section 2.1.2).
2.3. 1 Neutralization
Acidic wastes with a pH < 2.0 and alkaline wastes
with a pH >12.5 are defined as hazardous (40 CFR
Part 261). To meet the regulatory definition of non-
hazardous, acidic wastes must be neutralized to a
pH > 2.0 by reducing the hydrogen ion concentration,
and alkaline wastes must be neutralized to a pH >
12.5 by increasing the hydrogen ion concentra-
tion.
Carbonates (limestone and dolomite) will dissolve in
and neutralize acidic wastes. The process is:
CaCOs — > Ca+2 + COa"2 (dissolution)
CO3
"2
CO2 + H2O (neutralization)
When calcium carbonate goes into solution, it releases
basic carbonate ions (CDs'2), which react with hydrogen
tons to form carbon dioxide (which will normally remain in
solution at deep-well-injection pressures) and water.
Removal of hydrogen tons raises the pH of the solution.
However, aqueous carbon dioxide serves to buffer the
solution (i.e., re-forms carbonic acid in reaction with
water to add H+ ions to solution). Consequently, the
buffering capacity of the solution must be exceeded
before complete neutralization will take place. Buffer-
ing capacity and the specific chemical reaction in-
volving carbon dioxide, water, and carbonic acid are
discussed in more detail in Section 3.1.1. Nitric acid
can react with certain alcohols and ketones under in-
creased pressure to increase the pH of the solution,
21
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Table 2-6 Examples of the Effects of Transformation Processes on the Toxicity of Substances
Type of Transformation
Process
Source
Detoxification
Cyanide — > Amide — > Acids + Ammonia
Cyanide — > Sulfate + Carbon + Nitrogen
Nitrile — > Amide — > Acids + Ammonia
Alkyl halide — > Alcohol + Halide ion
Chlorobenzene — > CO2 + CI" + HaO
1 ,3-Dichlorobenzene — > CO2 + CI " + HaO
1 ,4-Dichlorobenzene — > CO2 + CI " + HteO
Vinyl chloride — > CC-2 + CI" + h^O
Hydrolysis
Biooxidation
Hydrolysis
Hydrolysis
Biooxidation
Biooxidation
Biooxidation
Bioreduction
Scrivneret al., 1986
Mudder and Whittlock
Scrivneret al., 1986
Valentine, 1986
Bouwer and McCarty,
ibid.
ibid.
, 1983
1984
Vogel and McCarty, 1987
Transtoxification
2,4-D ester —> 2,4-D acid (increased) Hydrolysis
Phenol + Formaldehyde —> Phenolic resins Polymerization
Aldrin —> Dieldrin Oxidation
DDT—> ODD Reduction
o-Xylene —> o-Toluic acid Co-metabolism
Benzene —> Phenol Biooxidation
Carbon tetrachloride —> Chloroform —> Bioreduction
Methylene chloride
Ethylbenzene —> Phenylacetic acid Co-metabolism
1,1,1-Trichloroethane—> Bioreduction
1,1-Dichloroethane —> Chloroethane
Tetrachloroethylene —> Bioreduction
Trichloroethylene —> Various
Dichloroethenes —> Vinyl chloride
1,2-Dichloroethane —> Vinyl chloride Hydrolysis
Inorganic mercury —> Methyl mercury Bioreduction
Nitrilotriacetate—> Nitrosamines Bioreduction
Mills etal., 1985
Strycker and Collins, 1987
Crosby, 1973
Glass, 1973
Horvath, 1972
Gibson, 1972
Wood et al., 1985
Horvath, 1972
Wood etal., 1985
ibid.
Ellington et al., 1988
Reederetal., 1977
ibid.
Toxification
Amines —> Nitrosamines
Biooxidation
Alexander, 1981
22
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and this reaction was proposed by Goolsby (1971) to
explain the lower-than-expected level of calcium ions
in backflowed waste at the Monsanto waste injection
facility in Florida (see Section 7.2).
Quartz (SiCte) and other silicates are generally stable
in acidic solutions but will dissolve in highly alkaline
waste solutions, decreasing the pH of the waste. The
process by which this reaction occurs is complicated
because it creates complex mixtures of nonionic and
ionic species of silica. Scrivner et al. (1986) discusses
these reactions in some detail. They observe that the
silicates in solution buffer the liquid. Also, laboratory
experiments in which alkaline wastes have been
mixed with sandstone have shown relatively small
reductions in pH. At near-surface temperature and
pressure conditions, an alkaline waste remains haz-
ardous, but at simulated subsurface temperatures
and pressures, the waste is rendered nonhazardous,
ranging in pH from 11.5 to 12.4 in the experiments
performed by Roy et al. (1989). However, the pH of
the sandstone-waste mixture remained above 12.5 in
other investigations, possibly because a higher
solid:liquid ratio (less sandstone per volume of liquid)
is used (personal communication, May 10, 1990,
W. R. Roy, Illinois State Geological Survey, Cham-
paign, Illinois).
Reactions with clay minerals can neutralize both low-
pH and high-pH solutions. Neutralization of acids oc-
curs when hydrogen ions replace Al, Mg, and Fe. In
alkaline solutions, neutralization is more complex
and may involve cation exchange, clay dissolution,
and reaction of cations with hydroxide ions to form
new minerals called zeolites (Scrivner et al., 1986).
2.3.2 Complexation
A complex ion is one that contains more than one
ion. Because of its effect on mobility, complexation,
the process by which complex ions form in solution,
is very important for heavy metals and may be sig-
nificant for organic wastes. Heavy metals are par-
ticularly prone to complexation because their atomic
structure (specifically the presence of unfilled dorbi-
tals) favors the formation of strong bonds with polar
molecules, such as water and ammonia (NHa), and
anions, such as chloride (Cl~) and cyanide (CN ~).
Depending on the chemistry of an injected waste
and existing conditions, complexation can increase
or decrease the waste's mobility.
Complexation is more likely in solutions with high
ionic strength (which is typical of fluids found in the
deep-well-injection environment—see Section 3.1.3).
This is true because the large number of ions
present in solution increases the number of chemical
species that can form (Langmuir, 1972). Many vari-
ables affect the stability of a complex ion relative to
ions and metals that can serve as potential ligands to
the central metal, the most important of which is the
valence (charge) of the central cation and its radius.
As a rule, the stability of complexes formed with a
given ligand increases with cation charge and
decreases with cation radius (Langmuir, 1979).
The solubility of most metals is much higher when
they exist as organometallic complexes (Strycker
and Collins, 1987). Naturally occurring chemicals
that can partially complex with metal compounds and
increase the solubility of the metal include aliphatic
acids, aromatic acids, alcohols, aldehydes, ketones,
amines, aromatic hydrocarbons, esters, ethers, and
phenols. Several complexation processes, including
chelation and hydration, can occur in the deep-well
environment.
2.3.2.1 Chelation
Chelation is the process of forming complex ions
with organic ligands that have more than one site
able to bond to the central metal ion in the complex.
The complex ion formed by this process is called a
chelate. The ligands in chelates are classified ac-
cording to the number of binding sites in the
molecule: monodentate (one site), bidentate (two
sites), etc. Metal solubility (i.e., mobility) is greatly in-
creased when chelation occurs, and metal-chelate
compounds are very stable when the metal ion is
chelated by a heterocyclic ring of an organic
molecule. Although most simple organic-metal com-
plexes will dissociate if solutions become more
dilute, chelated complexes do not tend to dissociate
(Martell, 1971). Even adsorbed metals may be remobil-
ized into solution by organic chelates. For example, the
synthetic chelate nitrilotriacetic acid (NTA), used as an
alternative to polyphosphate in detergents, has been
observed to remobilize adsorbed heavy metals in the
near-surface environment (Forstner and Wittmann,
1979). Although remobilization of heavy metals by
chelation has not been reported in the published litera-
ture on deep-well injection, the possibility should be
considered if the waste contains chelates.
2.3.2.2 Hydration
Metal ions in solution readily form complex ions by the
process of hydration (bonding to water molecules).
Because of the polar nature of water molecules, the
negative poles are attracted to the positively charged
metal ion, usually by ion-dipole bonding. Covalent
bonding may also occur (see Section 2.2.2 for discus-
sion of ion-dipole and covalent bonding).
23
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Hydration tends to increase the complexity of chemi-
cal reactions because hydrated polyvalent metal ions
may form multiple associations with other metals to
create complex polynuclear ions. Hydration may also
reduce mobility of metal ions through physical ad-
sorption, even when ion-exchange reactions are
completed in a solution. Reduction of mobility can
also be caused by dehydration when organic ligands
replace water molecules in complex ions. Polynuclear
metal ions and large organic complexes can be readily
adsorbed onto mineral surfaces because of their large
molecular weights, which enhance physical adsorption
(see Section 2.2.2). Metal complexes may also serve as
catalysts for a number of other chemical transformation
reactbns. These reactions are discussed in Section 2.3.5.
2.3.3 Hydrolysis
Hydrolysis occurs when a compound reacts chemi-
cally with water (i.e., new chemical species are
formed by the reaction), and can be a significant
transformation process for certain hazardous wastes
in the deep-well environment (see Table 2-7).
Hydrolysis reactions fall into two major categories:
replacement and addition. Each is discussed in
Section 2.3.3.1. The rates at which these reactions
occur are also significant in a fate assessment be-
cause some take so long to occur that they will not
take place during the analytical time frame (10,000
years).
2.3.3.1 Types of Hydrolysis
Replacement and addition are essentially irreversible
transformation processes. Both can be significant in
detoxifying some types of organic hazardous wastes.
In replacement, the most common hydrolysis reac-
tion, one functional group is replaced by an -OH
(hydroxide ion) originating from a water molecule.
For example, an hydroxide ion can replace the halide
ion in an alkyl halide to form an alcohol, leaving the
halide ion in solution. Alcohols can also form by addi-
tion of water to a carbon-carbon double bond.
Hydrolysis reactions can produce intermediate com-
pounds subject to further hydrolysis (e.g., nitriles to
amides to acids). Whether hydrolysis results in
detoxification, transtoxification, or toxification depends on
the toxicity of the most stable end product of any series
of hydrolysis reactions.
2.3.3.2 Hydrolysis Rates
Detoxification by hydrolysis is significant in fate as-
sessments only if the rate is fast enough to reduce
concentrations to acceptable levels at subsurface
locations of regulatory concern. Hydrolysis rates are
commonly reported in terms of half-life (i.e., the num-
ber of days or years for half of the original concentra-
Table 2-7 Listed Hazardous Organic Wastes for
which Hydrolysis May Be a Significant
Transformation Process in the Deep-Well
Environment
Group/Compound
Half-life*
Pesticides
DDT 81-4,400 b
Dieldrin 3,800
Endosulfan/Endosulfan sulfate 21
Heptachlor 1
Halogenated Aliphatic Hydrocarbons
Chloroethane (ethyl chloride) 38
1,2-Dichloropropane 180-700°
1,3-Dichloropropene ~60°
Hexachlorocyclopentadiene 14
Bromomethane (methyl bromide) 20
Bromodichloromethane 5,000
Halogenated Ethers
bis(Chloromethyl) ether <1
2-Chloroethyl vinyl ether 1,800
Monocyclic Aromatics
Pentachlorophenol 200
Phthalate Esters
Dimethyl phthalate 1,200
Diethyl phthalate 3,700
Di-n-butyl phthalate 7,600
Di-n-octyl phthalate 4,900d
aUnless otherwise indicated, half-life is measured in days
at pH = 7 and ambient temperature.
bCallahan et al. (1979) report lower value for pH 9 and
upper value for pH 3 to 5.
cEstimated values in Callahan et al. (1979); not based on
actual measurements.
dEllington et al. (1988) report a value of 107 years (39,000
days).
Sources: Callahan et al. (1979), Mills et al. (1985),
Schwarzenbach and Giger (1985), Ellington et al. (1988).
tion of the substance to be hydrolyzed). Predicted
hydrolysis half-lives of various hazardous organic
compounds range from days to thousands of years.
Such factors as pH, temperature, and the presence
of other ions affect the rate of hydrolysis of organic
compounds. Strycker and Collins (1987) speculate
that deep-well environments may lead to shorter
half-lives because of increased temperatures,
pressures, and Eh changes. Hydrolysis reaction
rates do increase with increasing temperatures, but
predicting rates in the deep-well environment is
24
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complicated because the influence of temperature
on hydrolysis is not always known precisely. The ef-
fects of ionic strength on hydrolysis reactions are
also difficult to predict and can lead to either ac-
celeration or retardation of rates, depending on the
substrate, the salts, and their concentrations (Mabey
and Mill, 1978). The high ionic concentrations of
reservoir brines in the deep-well environment make
this factor important in fate assessment.
A number of alkaline-earth and heavy-metal ions catalyze
hydrolysis of a variety of organic esters; however, this
reaction does not appear to be a major contributor to
hydrolysis under near-surface conditions (Mabey and Mill,
1978). Consequently, hydrolysis of susceptible com-
pounds may be enhanced in the deep-well environment
where concentrations of both alkali metal and heavy
metals in reservoir fluids and injected wastes are
much higher than those in typical near-surface environ-
ments.
Hydrolysis rates greatly depend on pH and vary
widely for an individual compound under acidic to
basic conditions. Chloromethane, which shows no
significant change in hydrolysis rate from a pH of 3 to
9, is an exception (Mills et al., 1985). Hydrogen
cyanide illustrates the strong effect that pH can have
on hydrolysis rates. Cyanides hydrolyze to amides,
which then hydrolyze to acids and ammonia. At pH>
10, this reaction has a half-life of about 10 years. At
pH 4, however, the reaction may take more than
10,000 years (Scrivner et al., 1986). Furthermore,
metal-cyanide complexes do not hydrolyze readily
and can reduce the concentration of free cyanide in
solution, increasing the time needed for the total
cyanide concentration to decrease from hydrolysis
(Scrivner et al., 1986).
Many classes of organic compounds hydrolyze in
aqueous solutions, whereas others are resistant.
Table 2-8 summarizes organic functional groups that
are potentially susceptible to hydrolysis and those
which are generally resistant. Only 8 out of 129
priority pollutants have half-lives on the order of 105
days or less in near-surface aquatic environments
(see Table 2-7).
At near-surface conditions, hydrolysis half-lives on
the order of hundreds of days may not be acceptable
(i.e., the reaction rate is too slow to reduce con-
centrations to standards). However, the 10,000-year
no-migration standard for deep-well-injected wastes
in EPA regulations (40 CFR 148.20) implies that half-
lives of hundreds and perhaps thousands of days
may result in significant reductions in waste con-
centrations before the waste has migrated sig-
Table 2-8 Amenability of Organic Functional Groups
to Hydrolysis
Potentially Susceptible Generally Resistant8
Alkyl halides
Amides
Amines
Carbamates
Carboxylic acid esters
Epoxides
Nitriles
Phosphonic acid esters
Phosphoric acid esters
Sulfonic acid esters
Sulfuric acid esters
Alkanes
Alkenes
Alkynes
Benzenes/biphenyls
Polycyclic aromatic
hydrocarbons
Heterocyclic polycyclic
aromatic hydrocarbons
Halogenated aromatics/PCBs
Dieldrin/Aldrin and related
halogenated hydrocarbon
pesticides
Aromatic nrtro compounds
Aromatic amines
Alcohols
Phenols
Glycols
Ethers
Aldehydes
Ketones
Carboxylic acids
Sulfonic acids
Multifunctional organic compounds in these categories
may be hydrolytically reactive if they contain other functional
group(s) that are hydrolyzable.
Source: Guswa et al. (1984), adapted from Harris (1982).
nificantly. Half-lives on the order of hundreds of days
(103) would go through at least 3,650 decay periods
in 10,000 years; half-lives on the order of magnitude
of thousands of days (104) would go through at least
365 periods in the same time. In other words, in the
course of 10,000 years, the concentration of a com-
pound with a half-life of 9,999 days would be reduced
by half, 365 times, which would almost certainly
reduce concentrations to below any laboratory detec-
tion limits. Table 2-7 lists 10 hazardous compounds
with hydrolysis half-lives on the order of hundreds to
thousands of days.
2.3.4 Oxidation-Reduction
Oxidation-reduction (redox) reactions involve the loss
of electrons and increase in oxidation number (oxida-
tion) by one substance or system with an associated
gain of electrons and decrease in oxidation number
(reduction) by another substance or system. Thus for
every oxidation there must be a reduction. The oxida-
tion number of an atom represents the hypothetical
25
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charge an atom would have if the ion or molecule
were to dissociate.
Because redox reactions involve the transfer of
electrons, the intensity of redox reactions is measured
by electrical potential differences, termed Eh (redox
potential). Highly oxidizing conditions will have an Eh of
about 0.8 volts; highly reducing conditions, an Eh of
about -0.4 volts. Eh as an environmental factor is dis-
cussed in Section 3.1 .2. Eh is difficult to measure ac-
curately, and ground-water systems are often not in
equilibrium with respect to redox reactions. Conse-
quently, the Eh of a chemical system indicates the
types of redox reactions that may occur rather than
predicting the specific reactions that are occuring. In in-
organic chemical systems, redox reactions tend to be
reversible, whereas microbblogically mediated redox
reactions involving hydrocarbons tend to be irre-
versible. Therefore, inorganic oxidation-reduction equi-
libria are somewhat analogous to acid-base equilibria.
2.3.4. 1 Redox Reactions Involving Simple
Hydrocarbons
The simplest oxidation reaction of an organic com-
pound is the transformation of a simple hydrocarbon
(such as a straight-chained compound) to carbon
dioxide and water in the presence of oxygen:
CH4 + 2O2
2H2O
This type of reaction is called aerobic respiration,
and without biological mediation it is irreversible.
As discussed in Section 2.2.1, acid-base reactions
change proportions of neutral and ionic species in
response to changes in pH (an analog of Eh) without
the intervention of transformation processes. In con-
trast, changes in Eh in ground-water systems
changes the type of oxidation-reduction that takes
place. Aerobic respiration quickly depletes dissolved
oxygen, and unless a continual supply of oxygen is
available, a sequence of reducing reactions is in-
itiated.
Table 2-9 shows the sequence of reducing reactions in-
volving formaldehyde that will occur after oxygen is
depleted in a closed ground-water system (i.e., when
there is no source of oxygen replenishment). Except in
unusual circumstances, when wastes contain sig-
nificant amounts of a strong oxidant such as chromium
(VI), reducing conditions will predominate in deep-well
injection zones. For example, conditions favoring
sulfate-reduction and methane-fermentation reactions
are most likely to occur in injection zones (see case
studies in Sections 7.2, 7.3, 7.4, and 7.5).
2.3.4.2 Redox Reactions Involving Complex
Organic Compounds
Oxidation reactions involving cyclic hydrocarbons and
hydrocarbon derivatives are more complex than those
for simple hydrocarbons, and it is not always obvious
how to classify such reactions in redox terms. In or-
ganic redox reactions, atoms, not electrons, usually are
transferred. Oxidation frequently involves a gain in
oxygen and a toss in hydrogen atoms, whereas
reduction involves the reverse. Organic functional
groups can be ranked by increasing oxidation state
Table 2-9 Redox Reactions in a Closed Ground-Water System
Reaction
Equation
1. Aerobic respiration
2. Denitrification
3. Mn(IV) reduction
4. Fe(lll) reduction
5. Sulfate reduction
6. Methane fermentation
7. Nitrogen fixation
CH2O + O2 = CO2 + H2O
5CH2O + Nitrate (4NO3>4H+ . Nitrogen (2N2) + 5CO2 + 7H2O
CH2O + 2MnO2 + 4H+ = 2Mn+2 + CO2 + 3H2O
CH2O + 8H+ + 4Fe(OH)3 = 4Fe+2 + CO2 + 11H2O
2CH2O + Sulfate (SO4~2) + H+ = HS" + 2CO2 + 2H2O
2CH2O + CO2 = Methane (CH4) + 2CO2
3CH2O + 3H2O + 2N2 + 4H+ = Ammonia (4NH4+) + 3CO2
Note: Reactions will tend to go to completion in sequence from top to bottom.
Source: Adapted from Champ et al. (1979).
26
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to facilitate classification of reactions as either oxida-
tion or reduction. Table 2-10 summarizes relative
oxidation states of several major functional groups. A
functional group is considered oxidized if it is con-
verted into a functional group at a higher oxidation
state. Reduction is conversion to a group at a lower
state.
Table 2-10 Relative Oxidation States of Organic
Functional Groups
Functional Group Oxidation State
-4 -2 0 +2 +4
Least Oxidized Most Oxidized
RH ROM RC(O)R RCOOH CO2
RCI (R)2CCI2 RC(0)NH2 CCI4
RNH2 RCCI3
C = C -C=C-
Source: Adapted by Valentine (1986) from March (1977).
Table 2-11 lists some organic compounds according
to their susceptibility to oxidation. Oxidation reactions
are more common in the near-surface environment,
where oxygen is abundant and sunlight may provide
additional energy for reactions to take place. Oxida-
tion is usually not a significant process in the deep-
well environment except, perhaps, when strong
oxidants such as Cr (VI) or permanganate are part of
the injected waste (Strycker and Collins, 1987).
In general, the importance of redox reactions involv-
ing organic compounds in soil and water is not well-
documented (Callahan et al., 1979; Valentine, 1986).
Table 2-11 Susceptibility of Organic Compounds to
Oxidation In Water
Most-Susceptible
Least-Suscept ib le
Phenols
Aromatic amines
Olefins and dienes (electron-rich)
Alkyl sulfides
Enamines
Alkenes
Haloalkanes
Alcohols
Esters
Ketones
Source: Mill (1980).
In anaerobic environments, typical of deep-well-
injection zones, reduction of chemicals by both
biological and nonbiological processes can occur.
Reduction of organochlorine compounds (such as
DDT and toxaphene), in which a chlorine atom is
replaced by a hydrogen atom, is the most frequently
reported example of this type of reaction (Callahan et
al., 1979).
2.3.5 Catalysis
The rates of many reactions increase in the presence of
a catalyst, which itself remains unchanged in quantity
and composition afterward. Although the catalyst itself is
not transformed, the catalyst speeds up reactions that
would occur naturally or promotes reactions that
would not occur otherwise. For example, metal ions
catalyze the hydrolysis and oxidation reactions in
biochemical systems (Martell, 1971). Phenol and
phenol derivatives are normally resistant to oxidation in
wastewaters, but the reaction can be accomplished
by metal-ion catalysis when Fe+2, Mn+2, Cu+2, and
Co"1"2 are combined with chelating agents (Martell,
1971). The reactions involved in destroying the
aromatic ring in these compounds are complex and
more likely to occur during waste pretreatment than
as a result of processes in the deep-well environ-
ment. Certain metals in the presence of clays can
also catalyze the polymerization of phenols and ben-
zenes (see Section 2.3.6). Laszlo (1987) reviews or-
ganic reactions that are catalyzed by clay minerals.
2.3.6 Polymerization
Polymerization is the formation of large molecules
(polymers) by the bonding together of many smaller
molecules. For example, styrene polymerizes to form
polystyrene. Polymerization can enhance the ten-
dency of a substance to be adsorbed on mineral sur-
faces by increasing the molecular weight but is not
likely to result in detoxification of hazardous wastes.
Polar organic compounds such as amino acids nor-
mally do not polymerize in water because of dipole-
dipole interactions. However, polymerization of
amino acids to peptides may occur on clay surfaces.
For example, Degens and Metheja (1971) found
kaolinite to serve as a catalyst for the polymerization
of amino acids to peptides.
Adsorption of phenol and benzene as a result of
polymerization at the clay surface has also been ob-
served in the laboratory on smectite clay (in the
montmorillonite group) when exchangeable sites
were occupied by Fe+3 or Cu+2 cations (Mortland
and Halloran, 1976). In natural systems, Cu+2 is not
very likely to exist in significant enough concentra-
tions. However, Fe"1"3 may be present in the deep-
27
-------
well environment in sufficient amounts to enhance
the adsorption of phenol, benzene, and related
aromatics. Wastes from resin manufacturing facilities,
food processing plants, pharmaceutical plants, and other
types of chemical plants occasionally contain resin-like
materials that may polymerize to form solids at deep-
well-injection pressures and temperatures (Selm and
Hulse, 1960).
2.3.7 Thermal Degradation
Thermal degradation occurs when heat causes
compounds to undergo structural changes, leading
to formation of simpler species. For example, many
organophosphorus esters isomerize when heated
and break down into component molecules (Crosby,
1973). Temperatures and pressures common in the
deep-well environment are normally too low to in-
itiate high-temperature reactions, but if the right
chemicals (not necessarily hazardous) are present,
thermal degradation might be initiated (Strycker and
Collins, 1987). For example, thermal decarboxylation
is probably the mechanism of acetate degradation in
oilfield waters where temperatures exceed 200°C
(Carothers and Kharaka, 1978; Kharaka et al.,
1983); however, injection zones usually do not reach
this temperature. At a depth of 900 meters (ap-
proximately 3,000 ft) temperatures range from 50° to
100°C(Royetal., 1989).
Smith and Raptis (1986) have suggested using the
deep-well environment as a wet-oxidation reactor for
liquid organic wastes. This process, however, does
not involve deep-well injection of wastes but rather
uses temperatures and pressures in the subsurface
to increase the oxidation rate of organic wastes,
which are then returned to the surface.
2.3.8 Biodegradation
Biotransformation is the alteration of a compound
as a result of the influence of organisms. It is one of
the most prevalent processes causing the break-
down of organic compounds in the near-surface en-
vironment. Biodegradation is a more specific term
used to describe the biologically mediated change of
a chemical into simpler products. The term includes,
and sometimes obscures, a series of distinctive
processes of toxicological significance in natural
ecosystems. Biodegradation is probably more sig-
nificant in the decomposition of the nonhazardous
components of deep-well-injected organic wastes
(see Case Studies, Sections 7.2, 7.3, 7.4 and 7.5),
although a few hazardous compounds, such as
acrylonitrile (see Case Study, Section 7.3) and some
monocyclic aromatic hydrocarbons and halogenated
aliphatics, may be subject to biodegradation in the
deep-well environment (see Section 3.4).
Microorganisms are by far the most significant group
of organisms involved in biodegradation (Scow,
1982). They can mineralize (convert to CC-2 and
HaO) many complex organic molecules that higher
organisms, such as vertebrates, cannot metabolize.
They are often the first agents in biodegradation,
converting compounds into the simpler forms re-
quired by higher organisms. Most biodegradation in
near-surface environments is carried out by
heterotrophic bacteria (microorganisms that require
organic matter for energy and oxygen).
Biodegradation in deep-well environments is per-
formed predominantly by anaerobic microorganisms,
which do not consume oxygen and are either
obligate (oxygen is toxic to the organism) or faculta-
tive (the organism can live with or without oxygen or
prefers a reducing environment). The two main types
of anaerobic bacteria, methanogenic (methane-
producing) and sulfate-reducing, do not degrade
the same compounds (Strycker and Collins, 1987).
The by-products of sulfate reduction are hydrogen
sulfide, carbon dioxide, and water (see Equation 5,
Table 2-9). Methanogenic bacteria produce methane
and carbon dioxide (see Equation 6, Table 2-9). The
extent to which either type proliferates is strongly in-
fluenced by pH. As a group, anaerobic organisms
are more sensitive and susceptible to inhibition than
aerobic bacteria (Scow, 1982). Typically, aerobic
degradation also is more efficient than anaerobic
degradation, and high temperatures are not as limit-
ing for aerobes as for anaerobes (Strycker and
Collins, 1987).
Alexander (1980) identifies six major kinds of bio-
degradation:
• Mineralization
• Co-metabolism
• Detoxification
• Transtoxification
• Activation
• Defusing
Table 2-12 describes each of these processes and
gives examples.
For several reasons, mineralization (decomposition to
inorganic constituents) is generally a more effective
form of biodegradation than co-metabolism (conversion
to another compound without using the original
compound for energy or growth). First, detoxification is
28
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Table 2-12 Summary Descriptions of the Major Types of Biological Transformation Processes
Process
Description
The complete conversion of an organic compound to inorganic constituents (water, carbon dioxide).
Generally results in complete detoxification unless one of the products is of environmental concern,
such as nitrates and sulfides under certain conditions.
Conversion of an organic compound to another organic compound without the microorganism
using the compound as a nutrient. Resulting compounds may be as toxic (DDT to DDE or ODD)
or less toxic (xylenes to toluic acid).
Conversion of a toxic organic compound to a nontoxic organic compound. The pesticide 2,4-D
can be detoxified microbially to 2,4-dichlorophenol.
Conversion of a toxic compound to another toxic compound with similar, increased, or
reduced toxicity.
Conversion of a nontoxic molecule to one that is toxic, or a molecule with low potency to one that
is more potent. Examples include the formation of the phenoxy herbicide 2,4-D from the
corresponding butyrate, formation of nitrosamines, and methylation of arsenicals to trimethylarsine.
Conversion of a compound capable of becoming hazardous to another nonhazardous compound
by circumventing the hazardous intermediate. This has been observed in the laboratory, but not
identified in the environment. An example is the direct formation of 2,4-dichlorophenol from the
corresponding butyrate of 2,4-D.
Sources: Alexander (1980) and Horvath (1972).
Mineralization
Co-metabolism
Detoxification
Transtoxification
Activation
Defusing
more likely to occur during mineralization. Second,
mineralizing populations will increase until the
compound is completely degraded, because they
use the compound as a source of energy. In
contrast, co-metabolized compounds tend to change
slowly, and the original compound and its reaction
products tend to remain in the environment because
the co-metabolized compounds are not used for energy.
These differences are illustrated in Figure 2-1.
Almost all the specific chemical reactions in
biodegradation can be classified as oxidation-
reduction, hydrolysis, or conjugation. Hydrolysis
and oxidation-reduction have been discussed in Sec-
tion 2.3.3 and Section 2.3.4. Conjugation involves
the addition of functional groups or a hydrocarbon
moiety to an organic molecule or inorganic species.
For example, conjugation occurs when microbial
processes transform inorganic mercury into dimethyl
mercury.
At least 26 oxidative, 7 reductive, and 14 hydrolytic
transformations of pesticides had been identified as
of 1975 (Goring et al., 1975). Detailed identification
and discussion of specific reactions can be found in
Alexander (1981) and Scow (1982). Section 3.4 dis-
cusses the effects of environmental factors on haz-
ardous waste biodegradation in the deep-well en-
vironment.
2.4 Transport Processes
Many factors and processes must be considered
when evaluating the movement of deep-well-injected
hazardous wastes. Most of these processes are
beyond the scope of this reference guide. Four fac-
tors, however, are relevant to geochemical charac-
teristics:
• Hydrodynamic dispersion
• Osmotic potential
• Particle migration
• Density and viscosity
2.4.1 Hydrodynamic Dispersion
Hydrodynamic dispersion refers to the net effect
of a variety of microscopic, macroscopic, and
regional conditions that affect the spread of a
solute front through an aquifer (Mills et al.,
1985). Quantifying the dispersion is important to
fate assessment because contaminants can
29
-------
Figure 2-1 Hypothetical Model for Population Changes and Metabolism of a Chemical
Modified by Mineralizing and Co-metabolizing Populations (Alexander, 1981).
Mineralizing population 3
Tlma •
move more rapidly through an aquifer by this
process than would be predicted by simple plugf low
(i.e., uniform movement of water through an
aquifer with a vertical front). In other words,
physical conditions (such as more-permeable
zones, where water can move more quickly) and
chemical processes (e.g., movement of dissolved
species at greater velocities than the water
moves by molecular diffusion) result in more
rapid movement of contaminants than would be
predicted by ground-water equations for physi-
cal flow, which must assume average values for
permeability.
Dispersion on the microscopic scale is caused by:
• Velocity variations resulting from variations in
pore geometry and the fact that water velocity is
higher in the center of a pore space than that for
water moving near the pore wall
• Molecular diffusion along concentration gradients
• Variations in fluid properties such as density and
viscosity (Section 2.4.4)
Dispersion on the macroscopic scale is caused by
variations in hydraulic conductivity and porosity,
which create irregularities in the seepage velocity
with consequent mixing of the solute. Finally, over
large distances, regional variations in hydrogeologic
units can affect the amount of dispersion. In
hydrogeologic modeling, the hydrodynamic disper-
sion coefficient (D) is often expressed as the sum of
a mechanical dispersion coefficient (Dm) and
molecular (Fickian) diffusion (D*).
In most instances, hydrodynamic dispersion is not
great enough to require detailed consideration in
hydrogeologic modeling for fate assessment of deep-
well-injected wastes. However, regional variations
(such as presence of an underground source of
drinking water [USDW] in the same aquifer as the
injection zone, as is the case in parts of Florida)
should be evaluated before a decision is made to ex-
clude it.
2.4.2 Osmotic Potential
Osmotic potential refers to the energy required to pull
water away from ions in solution that are attracted to
the polar water molecules. In the presence of a semi-
permeable membrane between two solutions, water
molecules will move through the membrane to the side
with the higher concentration. This property may be im-
portant to fate assessment because in the deep-well
environment, shales that serve as confining layers can
act as semipermeable membranes if the injected waste
significantly changes the solute concentrations
(Hanshaw, 1972). In laboratory experiments, Kharaka
(1973) found that retardation sequences across geologic
30
-------
membranes varied with the material, but that
monovalent and divalent cations generally followed
identical sequences: Li < Na < NHa < K < Rb < Cs and
Mg < Ca < Sr < Ba.
If osmotic effects are possible, several other effects
would need to be considered in a geochemical-
fate assessment, depending on whether the
solute concentration is increased or decreased.
If solute concentrations are increased, pressures
associated with injection would increase beyond
those predicted without osmotic effects. Also, the
movement of ions to the injection zone from the
aquifer with lower salinity (above the clay confining
layer) would increase the salinity above those levels
predicted by simple mixing of the reservoir fluid and
the injected wastes. This action could affect the
results of any geochemical modeling.
If solute concentrations are decreased, the remote
possibility exists that wastes would migrate through
the confining layer. For this to occur, solute con-
centrations above the confining layer would have to
be higher than those in the injection zone, and move-
ment, in any event, would be very slow. Since USDWs
have salinities less than 10,000 mg/L, compared
with typical salinities in injection zones of 20,000 to
70,000 mg/L (see Sections 3.1.3 and 3.2.2), even if
this process were to occur it would cause migration
only to overlying aquifers that are not USDWs.
2.4.3 Particle Migration
Particle migration can occur when the mixing of in-
compatible fluids mobilizes clays or very fine par-
ticles precipitate out of solution. This process is
most likely to occur when solutions with low con-
centrations of salts are mixed with reservoir fluids
containing high concentrations, or when highly
alkaline solutions dissolve silica and release fines.
This type of reaction is of concern primarily when it
occurs near the injection zone, because particle
migration can clog pores and drastically reduce
permeability. McDowell-Boyer et al. (1986) provide
a good review of the literature on subsurface par-
ticle migration. Particle migration as it affects well-
plugging is discussed further in Section 3.3.1.
It is possible for complex metals ions that are ad-
sorbed onto very small particles of clay to migrate
as metal-clay particles. Laboratory experiments
found that radioisotope-clay particles at a low
salinity were retained in a sand core, but passed
through it at a high salinity (Strycker and Collins,
1987). Clay-metal particles would not be expected
to travel long distances in deep-well reservoir
rocks because the pores would be too small.
Injection of highly acid or alkaline wastes has the
potential to dissolve some reservoir rock to create
channels that would allow more distant transport of
small particles. Table 2-13 summarizes the various
physical parameters that affect particle migration in
porous-media flow.
2.4.4 Density/Viscosity Differences
Wastes having a different density (weight per unit
volume) or viscosity (tendency to resist internal flow)
than the injection zone fluids will tend to concentrate in
the upper (lower density/viscosity) or lower (higher
density/viscosity) portions of the injection zone.
Sniegocki (1960) discusses the effects of viscosity and
temperature on flow when there is little difference in
density between the injected and formation waters.
Kaufman and McKenzie (1975) observed that the ap-
parent hydraulic conductivity of the Belle Glade injec-
tion zone in Florida increased about 2.5 times because
of temperature differences (see Section 7.4).
Frind (1982) examines the basic requirements for
the mathematical simulation of density-dependent
transport in ground wafer. Miller et al. (1986) describe a
density-driven flow model designed specifically for
evaluating the potential for upward migration of
deep-well-injected wastes.
2.5 Interaction of Partition, Transfor-
mation, and Transport Processes
The actual movement of a specific deep-well-
injected hazardous substance depends on the
types of processes that act on the waste and on
the ways in which different processes interact.
Figure 2-2, from McCarty et al. (1981), shows
the expected change in concentration over time of
a deep-well-injected organic compound in an ob-
servation well at an unspecified distance from the
original point of injection.
With only dispersion operating, low concentrations
are observed before the arrival of a fluid exhibiting
ideal plug flow, but dispersion also serves to delay
the time it takes for 100% of the initial concentra-
tion to be observed. Adsorption combined with dis-
persion delays the arrival of the compound, and
eventually the contaminant will reach full con-
centration when adsorption capacity is reached.
When biodegradation occurs, initial concentra-
tions might well be governed by dispersion
alone, until sufficient time has passed for an
acclimated bacterial population to establish it-
self and become large enough to change the
31
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Table 2-13 Physical Parameters Affecting Particle Migration in Porous-Media Flow
Parameter
Significance
Matrix
Porosity
Particle size for which 10%
of the matrix is smaller
than that size
Particle size for which 60%
of the matrix is smaller
than that size
Bulk density
Specific surface area
Grain shapes
Surface roughness of grains
Pore-diameter size and size
distribution
Surface charge of grains
Indicates voids; space available for retention of clogging material.
Termed the effective size for filter sands.
The ratio of the 60% size to the 10% size is an indicator of the uniformity.
For a given material, indicates the closeness of packing and propensity for material
movement under stress.
Relates to surface-active phenomena and adsorption rate.
Affects shape of pores and thus fluid-flow patterns.
Affects retention of suspension on the particle surface.
Propensity for entrapment or filtration of suspension.
Negatively charged surface grains will attract a suspended particle with a positive charge.
Fluid
Viscosity
Density
Velocity of flow
Pressure
Shear forces and fluid resistance to flow.
Mixing effects when different densities are involved; may affect direction and rate of flow.
Hydrodynamic forces on the medium and suspension.
Driving force moving the liquid and suspension into and through the medium.
Suspended Particles
Concentration (inflow, within Material available for inflow, retention, and through-flow.
medium, outflow)
Size
Shape
Electric charge
Ability to pass through pore openings.
Effect on retention or through-flow due to orientation.
Attraction or repulsion to medium or intermediate materials.
Source: Adapted from Signer (1973).
32
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Figure 2-2 Effects of Dispersion, Adsorption, and Biodegradation on the Time Change in
Concentration of an Organic Compound in an Aquifer Observation Well. Following the
Initiation of water injection into the aquifer at some distance away from the observation
well, C represents the observed concentration and Co the concentration in the injection
water (McCarty, et al., 1981).
Expected responses to a step change in concentration
_—
Dispersion
. Ideal
plug
' flow I
Sorption
and dispersion
Biodegradation
and dispersion
Biodegradation,
sorption,
and dispersion
Time relative to mean residence
time of water. f'r"H,o
organic concentration significantly. If this occurs,
the concentration would decrease and level out
at some minimum value. When adsorption acts
with biodegradation, the arrival of the con-
taminant is delayed, as with adsorption alone;
then the concentration of the contaminant rises
to a maximum level below that of the original
concentration and declines as biodegradation be-
comes active.
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Mortland, M. M., and L. J. Halloran. 1976. Polymeriza-
tion of Aromatic Molecules on Smectite. So/7 Sci. Soc.
Amer. J. 40:367-370.
Mudder, T. I., and J. L. Whitlock. 1983. Biological
Treatment of Cyanidation Wastewaters. In Proc. of
the 38th Industrial Waste Conference, Purdue
University, West Lafayette, Indiana. Butterworth Pub-
lishers, Boston, pp. 279-287.
Palmer, C. D., and R. L. Johnson. 1989. Physical
Processes Controlling the Transport of Non-Aqueous
Phase Liquids in the Subsurface. In Transport and
Fate of Contaminants in the Subsurface, Chapter 3.
EPA 625/4-89/019.
Rao, P. S. C., and J. M. Davidson. 1980. Estimation of
Pesticide Retention and Transformation Parameters
Required in Nonpoint Source Pollution Models. In En-
vironmental Impact of Nonpoint Source Pollution, M. R.
Overcash, and J. M. Davidson eds. Ann Arbor Science
Publishers, Ann Arbor, Michigan, pp. 23-67.
Reeder, L. R., et al. 1977. Review and Assessment
of Deep-Well Injection of Hazardous Wastes (4
Volumes). EPA 600/2-77-029a-d, NTIS PB 269 001
to 004.
Roy, W. R., I. G. Krapac, S. F. J. Chou, and R. A.
Griffin. 1987. Batch-Type Adsorption Procedures for
Estimating Soil Attenuation of Chemicals. Draft Tech-
nical Resource Document (TRD), EPA/530-SW-87-
006-F. NTIS PB87-146155. [The final TRD, entitled
Batch-Type Procedures for Estimating Soil Adsorp-
tion of Chemicals, is scheduled for publication in
1990]
Roy, W. R., S. C. Mravik, I. G. Krapac, D. R.
Dickerson, and R. A. Griffin. 1989. Geochemical
Interactions of Hazardous Wastes with Geological
Formations in Deep-Well Systems. Environmental
Geology Notes 130. Illinois State Geological Survey,
Champaign, Illinois. [An earlier version of this report
by the same title was published in 1988 by the
Hazardous Waste Research and Information
Center, Savoy, Illinois].
Rubin, J. R. 1983. Transport of Reacting Solutes in
Porous Media: Relation between Mathematical Na-
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Reactions. Water Resources Research 19:1231-
1252.
Schwarzenbach, R. P., and W. Giger. 1985. Behavior
and Fate of Halogenated Hydrocarbons in Ground
35
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Water. In Ground Water Quality, C. H. Ward,
W. Giger, and P. L. McCarty, eds. Wiley Interscience,
New York, pp. 446-471.
Scow, K. M. 1982. Rate of Biodegradation. In Hand-
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Lyman, W. F. Reehl, and D. H. Rosenblatt, eds. Mc-
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Deep Well Wet Oxidation of Liquid Organic Wastes.
In Proc. of the Int. Symp. on Subsurface Injection of
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Strycker, A., and A. G. Collins. 1987. State-of-the-Art
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Amer. Proc. 38:29-35.24
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Biotic Transformations of 1,1,1-Trichloroethane under
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493-511.
36
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CHAPTER THREE
MAJOR ENVIRONMENTAL FACTORS AFFECTING DEEP-WELL-INJECTION
GEOCHEMICAL PROCESSES
Environmental conditions determine in large part the
chemical reactions that will occur when waste is in-
jected. For example, precipitation-dissolution reac-
tions are strongly controlled by pH. Thus iron
oxides, which may be dissolved in acidic wastes,
may precipitate when injection-zone mixing in-
creases the pH of the waste. Similarly, redox poten-
tial (Eh) exerts a strong control on the type of
microbiological degradation of wastes.
The most variable and site-specific factor is the reser-
voir rock matrix. Geologic formations vary greatly in
chemical and physical properties depending on the
conditions under which they formed and the geologic
processes to which they have been subjected. Other
environmental factors such as pH, Eh, salinity,
temperature, and pressure fall within a relatively well-
defined range in the deep-well environment, and thus,
to a certain extent, restrict the chemical processes that
can be expected.
This chapter has four sections:
• Section 3.1 (Major Environmental Factors Influenc-
ing Geochemical-Fate Processes) discusses how
the environmental factors of pH, Eh, salinity, reser-
voir matn'x, temperature, and pressure affect chemi-
cal processes.
• Section 3.2 (Geochemical Characteristics of Deep-
Well-lnjection Zones) examines the typical range of
environmental conditions that occurs in the deep-well
environment, including lithology (Section 3.2.1) and
the geochemical characteristics of deep-well brines
(Section 3.2.2).
• Section 3.3 (Influence of Environment on Waste-
Reservoir Compatibility) examines the possible
impacts of environmental factors and chemical
processes on the operation of injection wells.
These impacts include well plugging (Section
3.3.1), well-casing/ confining-formation failure
(Section 3.3.2), and well blowout (Section 3.3.3).
• Section 3.4 (Influence of Deep-Well Environmen-
tal Factors on Biodegradation) examines in detail
the significance of conditions in the deep-well en-
vironment as they affect biodegradation of injected
wastes.
3.1 Major Environmental Factors
Influencing Geochemical-Fate Processes
The previous chapter examined the geochemical
processes that can occur in the deep-well environment.
The type and outcome of reactions that will actually
occur when a waste is injected, however, depend on its
chemical characteristics (discussed in Chapter Four)
and on injection-zone conditions. This chapter ex-
amines six major environmental factors that must be
considered. Four (pH, Eh, salinity, and reservoir matrix)
are chemical properties or measures of chemical
properties of a system. These four provide information
on what types of reactions may occur and how they
might be expected to proceed. The remaining two fac-
tors, temperature and pressure, are physical properties
of the system that primarily influence reaction rates.
The purpose of this section is to provide a basic under-
standing of what these environmental factors are, how
they are measured or observed, and how they may af-
fect chemical processes.
3.1.1 pH
The symbol pH stands for the negative logarithm of
the hydronium ion [H3O+] activity and is a convenient
way of expressing the very low concentrations of
HaO"1" that are present in aqueous solutions. In
chemical reactions, the symbol H+ is often used in-
stead of HaO+. Pure water has a pH of 7. Solutions
with pH < 7 are acidic, and those with pH > 7 are
basic. Acid-base reactions (see Section 2.2.1) deter-
mine the pH of a solution at equilibrium.
The pH of a system greatly influences what chemical
processes will occur in the deep-well environment.
37
-------
Directly or indirectly, pH also affects most of the
other environmental factors that are discussed in this
chapter. Table 3-1 summarizes the significance and
some major effects of changes in pH on chemical
processes and environmental factors in the deep-
well environment.
Very small changes in acidity greatly affect chemical
reactions and the form of chemical species in solution.
For example, the hydrolysis half-life of hydrogen cyanide
is greater than 100,000 years at pH 4 but drops to about
10 years at pH 9 (Scrivner et al., 1986). Figures 4-3,
4-4, and 4-5 in Chapter Four illustrate how pH in-
fluences the distribution of molecular and ionic
species of cadmium, lead, and mercury.
Buffer capacity is a measure of how much the pH
changes when a strong acid or base is added to a
solution. A highly buffered solution will show little
change; conversely, the pH of a solution with low
buffering capacity will change rapidly. Weak acids or
bases buffer a solution, and the higher their
Table 3-1 Effects of pH on Deep-Well Geochemical Processes and Other Environmental Factors
Process/Factor
Significance of pH
Partition Processes
Acid-base
Adsorption-desorption
Precipitation-dissolution
Complexation
Hydrolysis
Oxidation-reduction
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. Mobility of heavy metals is strongly influenced by pH.
Adsorption of some organics is also pH-dependent.
Strongly influences precipitation-dissolution reactions. Mixing of solutions with different pH
often results in precipitation reactions. See also reservoir matrix below.
Transformation Processes
Strongly influences positions of equilibria involving complex ions and metal-chelate
formation.
Strongly influences rates of hydrolysis. Hydrolysis of aliphatic and alkylic halides optimum at
neutral to basic conditions. (Strycker and Collins, 1987). Other hydrolysis reactions tend to
be faster at either high or low pH (Kreitler et al., 1988).
Redox systems generally become more reducing with increasing pH (ZoBell, 1946).
Environmental Factors
Biodegradation
Eh
Salinity
Reservoir matrix
Temperature
Pressure
In combination with Eh, pH 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 al., 1960). In reducing
conditions, pH strongly affects whether methanogenic or sulfate-reducing bacteria
predominate (Strycker and Collins, 1987).
Increasing pH generally lowers Eh.
pH-induced dissolution increases salinity; pH-induced precipitation decreases salinity.
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.
pH-driven exothermic (heat-releasing) reactions will increase fluid temperature; pH-driven
endothermic (heat-consuming) reactions will decrease fluid temperature.
Will not influence pressure unless pH-induced reactions result in a significant change in the
volume of reaction products.
38
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concentration in solution, the greater the buffering
capacity. Alkalinity (usually expressed in calcium
carbonate equivalents required to neutralize acid to a
specified pH) is a measure of the buffering capacity
of a solution (Hem, 1970).
Acid-base reactions of buffers act either to add or to
remove hydrogen ions to or from the solution so as to
maintain a nearly constant equilibrium concentration of
H+. For example, carbon dioxide acts as a buffer when
it dissolves in water to form carbonic acid, which dis-
sociates to carbonate and bicarbonate ions:
CCtyaq) + hfeO —> HzCOa' <—> HCOa + H+ <—> CO3'2 + 2H+
(carbonic (carbon- (bicarbon-
acid) ate) ate)
At equilibrium, the concentration of H+ will remain con-
stant. When a strong acid (represented by H+ or HA) is
introduced into solution, the concentration of H+ is in-
creased. The buffer compensates by reacting with the
excess H+ tons, moving the direction of the above
reaction to the left. By combining with bicarbonate and
carbonate ions to form the nonionic carbonic acid,
equilibrium is reestablished at a pH nearly the same as
that existing before. The buffer capacity in this case is
determined by the total concentration of carbonate and
bicarbonate tons. When no more carbonate or bicar-
bonate tons are available to combine with excess H+
ions, the buffer capacity has been exceeded and pH
will change dramatically upon addition of further acid.
3.1.2 Eh and Other Redox Indicators
The term Eh, which is the oxidation-reduction poten-
tial, (often referred to as redox potential), is an expres-
sion of the tendency of a reversible redox system to be
oxidized or reduced. It is especially significant in its in-
fluence on btodegradatton processes (see Section 3.4).
The energy of oxidation (electron-escaping tendency)
present in a reversible oxidation-reduction system (in
volts [V] or millivolts [mV]) is measured as the potential
difference between a standard hydrogen electrode and
the system being measured. Large positive values (up to
about +800 mV) indicate an oxidizing tendency, and
large negative values (down to about -500 mV) indicate
a strong reducing tendency. Eh values of +200 mV and
tower indicate reducing conditions in near-surface soils
and sediments (Ponnamperuma, 1972).
The Eh of connate waters (water entrapped in the inter-
stices of sediment at the time of deposition) ranges
from 0 to -200 mV (Baas-Becking et al., 1960). For ex-
ample, formation water from two monitoring wells in the
lower limestone of the Floridan aquifer near Pensacola
ranged from +23 to -32 mV (Goolsby, 1972; see case
study in Section 7.2), and formation fluids from a
Devonian limestone in Illinois used for injection at a
depth of about 3,200 ft had an Eh of -154 mV (Roy et
al., 1989).
Several measures of organic pollutant loading to
waters have been developed to indicate the redox
status of a system: (1) biochemical oxygen demand
(BOD), (2) chemical oxygen demand (COD), (3) total
organic carbon (TOC), (4) dissolved organic carbon
(DOC), and (4) suspended organic carbon (SOC).
When values for any of these parameters are high,
oxygen is rapidly depleted in ground waters and
reducing conditions will develop. BOD and COD
were designed to measure oxygen consumption
during the microbial degradation of municipal
sewage. They are only semiquantitative indicators of
organic loading because measurement procedures
for these parameters have no direct geochemical
significance (Hem, 1970). Malcolm and Leenheer
(1973) recommend the use of DOC and SOC, which
are independent of microbial effects, toxic sub-
stance, and variability with diverse organic con-
stituents. TOC, when measured as a single
parameter (rather than as the sum of DOC and
SOC), provides less information for geochemical in-
terpretation (Malcolm and Leenheer, 1973).
Reducing conditions predominate in the deep-well
environment for several reasons:
• No source of oxygen replenishment exists.
• Higher temperatures in the deep-well environment
are associated with decreases in Eh.
• Neutral to slightly alkaline water in the deep-well
environment favors lower Eh values.
Deep-well injection of wastes can change, at least tem-
porarily, the Eh of the injection zone. For example,
Ragone et al. (1973) observed a change from reducing
to oxidizing conditions when tertiary-treated sewage
(reclaimed water) was injected into the Magothy
aquifer, Long Island, New York, at a depth of 400 ft.
The reclaimed water had 6.6 mg/L dissolved oxygen
compared with no dissolved oxygen in the formation
water. On the other hand, the Eh of an acidic waste
dropped dramatically, from +800 mV to around +100
mV, when mixed with siltstone under conditions of low
oxygen and simulated deep-well temperature and
pressure (Roy et al., 1989). Similarly, the Eh of an
alkaline waste dropped from +600 mv to about +200
mV (Roy et al., 1989). All the case studies in Chapter
Seven with sufficient data to evaluate redox potential
had reducing conditions (Monsanto, Section 7.2;
39
-------
American Cyanamid, Section 7.3; Belle Glade, Sec-
tion 7.4; and Wilmington, Section 7.5). In each case,
reducing conditions were indicated by the inorganic
byproducts of anaerobic microbial degradation.
3.1.3 Salinity and Specific Conductance
Salinity is defined as the concentration of total dissolved
solids (IDS) in a solution, usually expressed in mg/L.
The IDS concentration in water is usually determined
from the weight of the dry residue remaining after
evaporation of the volatile portion of the original solution.
Ground water may be classified into four salinity classes
(Hem, 1970):
• Slightly saline (1,000 to 3,000 mg/L)
• Moderately saline (3,000 to 10,000 mg/L)
• Very saline (10,000 to 35,000 mg/L)
• Brine (more than 35,000 mg/L)
Seawater is about 35,000 mg/L.
Water with a salinity of less than 10,000 mg/L is con-
sidered to be a potential underground source of drink-
ing water. By regulatory definition, deep-well injection
of hazardous waste can occur only in very saline
waters or brines. Actual salinities of waters in currently
used deep-well injection zones vary greatly (see Sec-
tion 3.2.2). In this reference guide, the term brines is
used to refer to the natural waters in deep-well injection
zones. As noted above, however, this term is not tech-
nically correct if IDS levels are less than 35,000 mg/L.
Solutions of substances that are good conductors of
electricity are called electrolytes. Sodium chbride, the
major constituent of seawater, is a strong electrolyte.
Most salts, as well as strong acids and bases, are
strong electrolytes because they remain in solution
primarily in ionic (charged) forms. Weak acids and
bases are weak electrolytes because they tend to
remain in nonionic forms. Pure water is a nonconductor
of electricity.
The conductivity of solutions is measured as specific
conductance, which may be expressed as micromhos
per centimeter (mhos/cm) or millimhos per centimeter
(mmhos/cm) at 25°C. Seawater has a specific conduc-
tance of about 50 mmhos/cm. Salinity shows a high
correlation with specific conductance at low to
moderate TDS levels, but the concentrations of ions in
brines are so high that the relationship between con-
centration and conductance becomes ill-defined (Hem,
1970).
As discussed in detail in Section 3.2.2 (Brine
Chemistry), in situ waters in the deep-well environment
have high salinities and also act as strong electrolytes.
Geochemical systems with these characteristics are
much more complex and difficult to model than those
with tow TDS levels (see discussion of Aqueous and
Solution Geochemistry Models, Section 5.2.1).
3.1.4 Reservoir Matrix
With few, if any, exceptions, deep-well injection zones
will be sedimentary rock, and the reactions that take
place when hazardous wastes are injected are deter-
mined largely by the physical and chemical properties
of that rock. The most important physical properties of
sedimentary rocks in relation to deep-well geochemical
interactions are texture (the proportions of different
sized particles in a sediment) and specific surface
area (see Section 6.2.2). The most important chemical
property is mineralogy, defined by the types and
proportions of minerals present.
3.1.4.1 Classification of Sedimentary Rocks
Sedimentary rocks can be broadly classified as clastic
and nonclastic. Clastic sediments include sandstones,
siltstones, and shales and are formed by the deposition
and cementation of soil and rock material that has
eroded from another location. These sediments are
classified primarily by particle size. The U.S. Depart-
ment of Agriculture (USDA) classification system is
commonly used to describe the size of grains in a
clastic sediment:
• Clay (<0.002 mm)
• Silt (0.002 to 0.05 mm)
• Sand (0.05 to 2.0 mm)
• Gravel (> 2.0 mm)
Clastic sedimentary rocks are classified according to
the predominant particle size: clay (shale), silt
(siltstone), or sand (sandstone). Note that the word clay
used in reference to texture has a distinctly different
meaning from its use in mineralogy. Clay particle size
is used to describe both clays and other minerals
less than 0.002 mm in diameter. Clay minerals are
classified according to crystalline structure (see
Section 3.1.4.2) and these minerals are typically,
but not always, in the clay-particle size range (see
Section 3.1.4.4)
Nonclastic rocks, in which precipitation (commonly car-
bonates) is the main contributor to rock formation, are
classified according to mineral composition. The most
important nonclastic rocks for deep-well injection are
40
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limestone (made up mostly of calcium carbonate)
and dolomite (calcium-magnesium carbonate). Non-
clastic rock texture can vary widely depending on the
proportions of clastic and precipitated material and
the extent to which dissolution and reprecipitation
occur after the sediment is deposited.
3.1.4.2 Sedimentary-Rock Minerals
A mineral is defined as a naturally occurring,
homogeneous solid with a definite chemical composition
and (usually) a well-defined crystalline structure. There
are hundreds of different minerals, but relatively few ac-
count for most of the volume in sedimentary rocks. In
sandstones and siltstones, quartz (silica, SiOa) is the
most common mineral, generally followed by feldspars
(potassium-, sodium-, or calcium-aluminum silicates). In
shales, a variety of clay minerals dominate. In limestone,
cateite (calcium carbonate) is the most common mineral.
The mineral dolomite (calcium-magnesium carbonate)
gives its name to strata composed of that mineral.
The rest of this section focuses on clay minerals be-
cause of their significance in defining adsorption
capacity, and also because of their possible contribu-
tion to well plugging (see Section 3.3.1). The impor-
tance of clays in catalyzing other reactions was
discussed in Section 2.3.5. Two broad groups are
recognized: silicate clays and hydrous-oxide
clays. Each is discussed below, along with a few ad-
ditional minerals.
Silicate Clays. Silicate clays have a sheetlike lattice
structure with either silicon (Si) in coordination with four
oxygen atoms (silica tetrahedra) or aluminum (Al) in
coordination with six oxygen (alumina octahedra). The
strong adsorptive capacity of clay is derived from the
negative charges created at the edges of these crystal-
line sheets, where oxygen atoms (O~2) have extra
electrons that are not bonded to the cations in the crys-
talline structure. The negative charge can be further in-
creased when ions with a tower valence substitute for
ions with a higher valence in the sheet structure (for ex-
ample, Al+3 substitutes for Si+4 in tetrahedral sheets
and Mg"1"2 for Al+3 in octahedral sheets).
Silicate clays are classified according to stacking
arrangements of the tetrahedral (silica) and octahedral
(alumina) lattice layers and their tendency to expand in
water. The stacking type strongly affects certain
properties, including (1) surface area, (2) the tendency
to swell during hydration, and (3) cation-exchange
capacity (CEC), the ability of a mineral surface to ad-
sorb ions. CEC is the sum of exchangeable cations
that a material can adsorb at a specific pH. It is com-
monly reported as milliequivalents per 100 grams
(meq), where 1 meq is 1 mg of hydrogen or the amount
of any other ton that will combine with or displace
1 mg hydrogen. The current Standard International
unit for reporting CEC is centimoles per kilogram
(Soil Science Society of America, 1987).
Table 3-2 summarizes some properties of silicate clay
minerals. The montmoriltonite group is most sensitive to
swelling and has a high CEC, because the 2:1 lattice
structure (two octahedral sheets separated by a
tetrahedral sheet) forms sheets that are toosely con-
nected by exchangeable cations. The exchange sites
between 2:1 lattice layers can be easily hydrated (i.e.,
adsorb water molecules) under certain conditions. Be-
cause the water molecules have a greater diameter than
the cations that hold the sheets together, hydration
pushes the layers apart. This process is discussed fur-
ther in Section 3.3.1 (Well Plugging). Vermiculite has
stronger negative charges on its inner surfaces than
montmorillonite because of the substitution of
lower-valence magnesium tons for aluminum. This fac-
tor results in an even higher CEC than that found in
montmoriltonite, but it also has the effect of bonding the
2:1 sheets more strongly. Consequently, vermiculite
clays are less susceptible to swelling.
In Table 3-2, the clays are listed from most reactive
(montmorillonite and vermiculite) to least reactive
(kaolinite). The 1:1 lattice structure in kaolinite creates
strong bonds between the paired sheets, resulting in a
low surface area and CEC. Illite and chlorite have inter-
mediate surface areas, CEC, and sensitivities to swell-
ing.
Clay minerals in sedimentary formations commonly
have characteristics of more than one clay mineral,
called mixed-layer clays. These minerals have
properties and compositions that are intermediate be-
tween two well-defined clay types (e.g., chlorite-
illite, illite-montmorillonite, etc.).
Hydrous-Oxide Clays. Hydrous-oxide clays are less
well understood than silicate clays (Brady, 1974).
These clays are oxides of iron, magnesium, and
aluminum associated with water molecules, although
the mechanism by which the water molecules are
held together is somewhat uncertain. Because of the
lower overall valence of the cations in hydrous-oxide
clays compared with silicate clays, CEC is lower.
Jenne (1968) suggests that hydrous oxides of mag-
nesium and iron furnish the principal control on the
fixation of cobalt, nickel, copper, and zinc heavy
metals in soils and freshwater sediments. Precipitation
of hydrous-oxide clays may also be significant in
waste-brine interactions, as discussed in Section 3.3.1.
41
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Table 3-2 Important Characteristics of Silicate Clay Minerals
Type of Clay"
Property
Montmorillonite Vermiculite
(Smectite)b
Illite
Chlorite
Kaolinite
Lattice type0
Expanding?
Specific surf ace
area (m2/g)
External surface
area
Internal surface
area
2:1
Yes
700-800
High
Very high
2:1
Slightly
700-800
High
High
2:1
No
65-120
Medium
Medium
2:2
No
25-40
Medium
Medium
1:1
No
7-30
Low
None
Swelling capacity
Cation exchange
capacity (meq/100 g)
Other similar
clays
High
80-120
Beidellite
Nontronite
Saponite
Bentonited
Med-High
100-150+
Medium
10-406
Low
10-406
Low
3-1 5e
Halloysite
Anauxite
Dickite
Nacrite
^Clays are arranged from most reactive (montmorillonite) to least reactive (kaolinite).
The term smectite is now used to refer to the montmorillonite group of clays (Soil Science Society of America, 1987).
c(tetrahedral:octahedral layers).
Bentonite is a clay formed from weathering of volcanic ash and is made up mostly of montmorillonite and beidellite.
"Upper range occurs with smaller particle size.
Sources: Adapted from Grim (1968), Brady (1974), and Ahlrichs (1972).
Other Minerals. Quartz and feldspar, the dominant sili-
cate minerals in clastic sedimentary formations in
terms of volume, also have negative charges on crystal
surfaces, which serve as exchange sites in adsorption-
desorption processes. The CEC of quartz and
feldspars is much tower than that for silicates in clays,
primarily because these minerals are found mostly in
the sand and silt fractions in sediments. However,
when waste is injected into sandstones, quartz and
feldspar can provide most of the exchange sites for ad-
sorption of waste constituents because they are a large
percentage of the rock by weight unless silicate clays
coat sand grains and line pore spaces or fractures (see
Section 3.1.4.4).
3.1.4.3 Organic Matter in Sedimentary Rocks
Forms of organic matter called humic substances
may significantly affect geochemical fate processes
in the deep-well environment, although this topic has
received little attention in this context. Important
chemical properties of humic substances include:
High adsorption capacity for metals and organic
pollutants (see Section 5.2.2.1).
Ready ability to form complexes with heavy
metals (Khan, 1980; Raspor et al. 1984; Weber,
1983).
Possible incorporation of organic pollutants that
have similar structures to the building block of
humus (such as chlorinated phenols, naphtholic
compounds, and halogenated anilines) into the
structure of humus during its formation (Bollag,
1983).
Ability to solubilize organic compounds that are
otherwise water-insoluble (Khan, 1980).
Ability to increase hydrolysis reactions as a catalyst
or, conversely, slow the rate of hydrolysis reactions
by adsorption (Perdue, 1983).
42
-------
• Potential to store both oxidizing and reducing
agents (Valentine, 1986).
• Ability to slow rates of other oxidation-reduction
reactions through adsorption (Valentine, 1986)
• Ability to strongly influence the size of microor-
ganism populations indigenous to deep-well for-
mations based on the amount of dissolved
organic matter in ground water.
Humic substances comprise a general class of
biogenic, refractory, yellow-black organic substances.
They are ubiquitous, occurring in all terrestrial and
aquatic environments. Although they have been studied
by scientists for about 200 years, no fundamental, or
even generally accepted, understanding of the nature,
origin, and geochemical role of humic substances has
been developed (Aiken et al., 1985).
Humic substances are classified into three major
groups: (1) humic acids (insoluble at pH 2; soluble at
higher pH), (2) fulvic acids (soluble under all pH con-
ditions), and (3) humins (insoluble residue). All humic
acids have colloidal properties. Their structure is
based primarily on six-membered aromatic and
heterocyclic rings, and may include benzene, naph-
thalene, and anthracene rings in their structure
(Manaskaya and Drozdova, 1968).
Humic Substances in Sedimentary Rocks.
Sedimentation removes organic materials from biologi-
cally active oxidizing environments with moderate
temperatures. When buried, organic materials are sub-
jected to reducing environments with high tempera-
tures and pressures that favor abiotic alterations of
these materials (Meinschein, 1971).
Organic matter in sedimentary rocks may be grossly
classified as bitumen, comprising organic substances
that are extracted by neutral organic solvents, and
kerogen, consisting of organic materials that are not
readily soluble in such solvents (McNabb and Dunlap,
1975). The bitumen fraction usually contains numerous
hydrocarbon materials, fatty acids, porphyrins and
many other substances, depending to some extent on
the organic solvent used for extraction. The kerogen
fraction constitutes the bulk of organic matter in subsur-
face environments, usually comprising 80 to 99 percent
of the total organic content of nonpetroleum-bearing
sedimentary materials (Tissot and Welte, 1978). The
composition of the kerogen fraction is very complex
and variable, generally consisting primarily of
humus-like materials.
Kerogen and other humic substances have been
studied most widely in relation to the formation of
petroleum deposits (Abelson, 1978; Tissot and
Welte, 1978; Durand, 1980). Kerogen may have a
wide variety of nitrogen, oxygen or sulfur (NOS)
functional groups attached to its surface, such as
-SOaH, -NH2, -COOH, -OH, as well as saturated
hydrocarbons and aromatic rings; NOS functional
groups in particular may be reactive adsorbents (Apps,
1988). The specific surface area of organic detritus in
near-surface sediments can be large and can make a
significant contribution to its total adsorption capacity
(Sposito, 1984; Karickhoff, 1984). On the other hand,
Grim (1968) states that organic matter in ancient sedi-
ments is not likely to exhibit as high a CEC as that in
near-surface sediments, because of metamorphosis.
The effect of humic substances on deep-well injection
merits additional research.
Organic Matter in Ground Water. Dissolved humic
substances in ground water may be geochemically sig-
nificant as complexing agents and as a substrate for
microorganisms. The low dissolved-organic-carbon con-
tent of most ground waters means that complexation of
heavy metals by humic substances generally will not be
a major process (Thurman, 1985). The amounts of dis-
solved organic matter in ground waters are sufficient to
support small but diverse populations of microorganisms
that may be able to adapt and degrade deep-well-
injected wastes (see Section 3.4).
Except as noted below, most ground waters contain less
than 1 mg dissolved organic carbon/liter. Kuznetsov et
al. (1963) report that dissolved organic matter in ground
waters typically ranges from tenths to tens of mg/L.
Thurman (1985), using data primarily from Leenheer et
al. (1974), reports the following median concentration of
organic carbon in various types of aquifers: sand and
gravel, 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 waters—
100.0 mg/L.
The origin of soluble organic compounds dissolved in
ground water is not well-understood, but it is generally
thought to consist mostly of humic substances, naph-
thenic acids, and phenolic compounds derived from the
organic matter in sedimentary rocks or lignin degrada
lion products derived from plant residue in surface soil
(Davis, 1967; McNabb and Dunlap, 1975; Matthess,
1982). Humic substances in deep aquifers appear to
have been derived primarily from kerogen associated
with sediments of the aquifer (Thurman, 1985). Meanj
(1982), in a study of the organic geochemistry of
ground water from deep aquifers near Hanford,
Washington (3,690 to 3,720 ft), and the Finnsjonn
(260-582 m), Sterno (320 m), and Stripa (350 m) mine
areas of Sweden, found that organic constituents in all
43
-------
the samples consisted predominantly of low-
molecular-weight fulvic acid.
Compared to surface waters, humic substances in
deep ground waters (greater than 150 m) typically
exhibit the following characteristics (Thurman, 1985):
• They account for less than a third of the dis-
solved organic carbon, compared to about 50%
for surface waters.
• They are more aliphatic (see Section 4.3), and
contain more carbon and less oxygen.
• They are similar to surface water in carboxyl con-
tent (5-6 meq/g), in molecular weights (1000-
5000) and in binding constants for copper (logK =
5.6 at pH 6.3).
As with the solid fraction, the dissolved fraction of
humic substances has received little study in the
context of deep-well injection. Data on dissolved or-
ganic carbon (Wilmington, North Carolina, and Pen-
sacola, Florida) and organic carbon (Belle Glade,
Florida) reported in Chapter Six, Table 6-6, are con-
sistent with typical values for sedimentary aquifers
reported above.
3.1.4.4 Relationship between Mineralogy and
Particle-Size Distribution
Table 3-3 shows the relationship between mineral-
ogy and particle size distribution in core samples of
the Frio formation, one of the most widely used
deep-well injection zones in Texas. Clay minerals,
particularly kaolinite, commonly occur as silt-sized
particles in this formation. Quartz is present mostly
as sand-sized particles, but in individual cores,
quartz may be as much as 48% silt-sized and 24%
clay-sized. Small percentages of feldspar (2% to
11%) and calcite (2% to 7%) may be in the clay size
class.
Particle-size distribution has a critical impact on
CEC, as shown in Table 3-4. The CEC of sand is
only 3.6 meq/100 g, compared with 32.8 for silt and
80.5 for clay (which includes clay-sized quartz and
feldspar particles). Although the silt and clay make
up 2.7% of the sediment by weight, they account for
27.1% of the CEC, with silt contributing about half
that. Nevertheless, even though sand has a very low
CEC, its large percentage by weight ensures that
most of the exchangeable sites (72.9%) are in that
fraction.
The mineralogical composition of pore surfaces and
intergrain voids that come in contact with injected
fluids usually is not the same as that of the bulk
Table 3-3 Mineral Composition and Particle-Size
Distribution of Core Samples of Upper
Frio Formation, Texas
Range of Percentage of Particle Size
Mineral
Sand
Silt
Clay
Montmorillonite
Illite
Kaolinite
Quartz
Feldspar
Calcite
0
0
2-6
69-93
2-27
0.3-6.4
0-7
1-5
23-42
18-48
2-8
2-11
1-12
1-9
39-75
11-24
2-11
2-7
Source: Kent (1981).
mass (Roy et al., 1989). Sandstone and siltstone
elastics, in particular, often have grain and pore sur-
faces coated with clays such as chlorite, illite, or
kaolinite. Consequently, such clays may be the
primary surface reacting with injected fluids even
though they may represent a small fraction of the
bulk mineralogy of the rock (Wilson and Pittman,
1977).
Particle size also affects the rate of decomposition
of organic matter by microorganisms. Messineva
(1962), in studies on the geological activity of bac-
teria in the Soviet Union, found that bacterial
mineralization of organic matter occurs most rapidly
in sand-silt sediments. In clay and clay-silt sediments
the process of mineralization is slowed, despite the
fact that the number of bacteria in the clay sediments
is considerably greater than that in sand-silt sedi-
ments. Sinclair and Ghiorse (1987) find similar
relationships between microbiological activity and the
saturated zone in near-surface aquifers: gravelly
sand is the most biologically active and clayey layers
the least.
3.1.5 Temperature and Pressure
Temperature and pressure are primary influences on
the rate of chemical reactions. Temperature is
measured in degrees using three main temperature
scales: Fahrenheit (°F), Centigrade (°C), and Kelvin
(°K). Pressure is measured in a variety of units, the
most common being atmospheres (atm), megapas-
cals (MPa), pounds per square inch (psi), and bars
(approximately equal to atmospheres). Both temperature
and pressure increase with depth below the earth's sur-
face. Consequently, temperatures and pressures in the
44
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Table 3-4 Effect of Particle Size on Cation-Exchange Capacity (CEC) of Natural Streambed Sediments, San
Mateo County, California
Material
Gravel
Sand
Silt
Clay
Diameter
(mm)a
<2.0
0.074-2.0
0.004-0.074
<0.004
CEC
(meq/100g)
0
3.6
32.8
80.5
Percentage
by Weight
2.6
94.7
1.9
0.8
Percentage
by CEC
0
72.9
13.3
13.8
aUSDA particle-classification boundaries differ slightly for silt-sand (0.05 mm) and clay-silt (0.002 mm).
Source: Brown (1979).
deep-well environment are significantly higher than
those in the near-surface environment.
Geotherma! gradients in the subsurface typically range
from 1°C per 50 ft to 1°C per 150 ft, with most regions
having a gradient of around 1°C per 100 ft (Roedder,
1959). Tables giving data on temperature gradients for
679 wells located in 23 states (Alabama, Arkansas,
California, Colorado, Illinois, Iowa, Kansas, Kentucky,
Louisiana, Michigan, Mississippi, Montana, Nevada,
New Jersey, New Mexico, North Dakota, Oklahoma,
Oregon, Pennsylvania, Texas, Washington, West Vir-
ginia, and Wyoming) can be found in Van Orstrand
(1934). Table 3-5, drawn from a variety of sources,
shows temperatures and pressures at various depths.
Temperature can vary greatly at the same depth in dif-
ferent locations. For example, temperatures at ap-
proximately the same depth in Florida differ by almost
26°C.
The velocity of most acid-base and dissolution reac-
tions increases as temperature increases. Higher
temperatures generally increase the rate of redox reac-
tions as well; however, the effect is difficult to predict
exactly because the interactions among competing
reactions may offset the effect of increased tempera-
ture (Valentine, 1986). In contrast, higher temperatures
usually decrease the amount and rate of adsorption,
because these reactions are generally exothermic
(heat-producing) (Strycker and Collins, 1987). An ex-
ception has been noted by Choi and Aomine (1974),
who found that adsorption rates of pentachlorophenol
on soil increase 6% to 12% when samples of three dif-
ferent soils are subjected to an increase in temperature
from 4° to 33°C. Adsorption decreased by 9% in a
fourth sample. Laboratory adsorption experiments at
constant, simulated deep-well pressure with phenol and
1,2-dichloroethane result in decreased adsorption with
increased temperature (Donaldson et al., 1975; Collins
and Crocker, 1988).
Greater pressures tend to decrease the growth and
survival of bacteria, but for certain species increased
temperature counters this effect. For example,
growth and reproduction of E. coll essentially stops
in nutrient cultures at 20°C and 400 atm (40.5 MPa).
When the temperature is increased to 40°C,
however, growth and reproduction are about the
same as at near-surface conditions (ZoBell and
Johnson, 1949).
Roy et al. (1989) conducted one of the most comprehen-
sive evaluations of geochemical interactions between
wastes and different rock types at elevated tempera-
tures and pressures. Their studies employed batch-
type tests in which two waste streams (acidic and
alkaline) were mixed separately with three rock forma-
tions (Mt. Simon sandstone, Proviso siltstone, and
Potosi dolomite) in short-term (15-day) studies. Long-
term (150-230/day) studies have been completed, but
the results have not yet been published. The short-
term studies were conducted at three temperature/
pressure combinations representing conditions at
the surface (25°C/0.1 MPa), at 1,500 ft (40°C/6.0
MPa), and at 3,000 ft (55°C/11.7 MPa). The long-term
studies were conducted at 52°C/10.8 MPa. Table 3-6
summarizes the effects observed in the short-term
studies of increasing temperature and pressure on pH,
Eh, and the concentrations of calcium, magnesium,
aluminum, silicon, and sulfate in solution. As tempera-
ture and pressure increase, neutralization generally
increases, with the greatest effect occurring in waste-
sandstone reactions and the least occurring in waste-
dolomite reactions. As noted in Section 2.3.1, Roy et
al. (1989) observed that increased temperature and
pressure are required to reduce the pH of alkaline
45
-------
waste below the regulatory limit of 12.5. In this study,
both the acid and alkaline systems show greater
reductions in Eh with increased temperature and
pressure, with the alkaline system showing the most
dramatic reductions. Table 3-6 indicates that higher
temperature and pressure both increase and lower
concentrations of Ca, Mg, Al, Si, and sulfate in solu-
tion, depending on the ion and the waste system.
Table 3-5 Temperature
Location/Depth
Feet Meters
and Pressure at Different Depths
Temperature Pressure
°C °F °K Atm MPa Psi
Sources
Illinois
5,300a 1,615
46 115a 319 155 15.7 2,275a
Kamath and
Salazar, 1986
Florida6
2,800a 850
2,900a 885
42 108a 315 — — —
16 61a 289 — — —
Henry and
Kahout, 1972
New York
4,050 1 ,235a
40a 104 313 122 12.4a 1,800
Ragone et al.,
1978
Texas/Tennessee
— —
60 140a 333 100a 10.1 1,470
Scrivneret al.,
1986
North Carolina
1 ,000a 305
75a 167 348 — — —
Peek and Heath,
1973
Texas0
3,499a 1 ,065
9,950a 3,035
3,450 1 ,050a
49a 120 322 — — —
1073 225 380 — — —
64a 147 337 112a 11.3 1,645
Kreitler et. al,
1988
Donaldson and
Johansen, 1973
Unspecified"
3,280 1 ,000a
50a 122 323 1323 13.4 1,940
— — — 205 20.73 3,000
Roedder, 1959
Collins and
Crocker, 1988
aValue(s) reported in citation.
bFrom two locations.
cFr\o formation (Kreitler et al., 1988) and Miocene sand (Donaldson and Johansen, 1973).
Typical pressure.
46
-------
Table 3-6 Effects of
Parameter
pH (Neutralization)
Acidic waste
Alkaline waste
Eh (Reduction)
Acidic waste
Alkaline waste
Ca Concentration
Acidic waste
Alkaline waste
Mg Concentration
Acidic waste
Alkaline waste
Al Concentration
Acidic waste
Alkaline waste
Si Concentration
Acidic waste
Alkaline waste
SCV2 Concentration
Acidic waste
Alkaline waste
Increased Temperature and
Mt. Simon
Sandstone
Greater
Slightly greater
Slightly greater
Much greater
Little change
Lowerb
Slightly higher
c
Lower
Higher
Higher
Higher
Slightly lower
Higher
Pressure on Waste- Rock Mixtures3
Proviso Siltstone
No correlation
Somewhat greater
Somewhat greater
Much greater
Slightly higher
Lower
Higher
c
Lower
Higher
No correlation
Higher
Lower
Slightly higher
Potosi Dolomite
No correlation
Somewhat greater
Somewhat greater
Little change
Higher
Lower0
Higher
c
Lower
Lower
No correlation
Lower
Higher
Higher
Parameter at higher temperature and pressure (55°C/11.7 MPa) compared to near surface conditions (25°C/0.1 MPa).
bPrecipitated.
cBelow analytical detection limit (0.07 mg/L).
Source: Adapted from Roy et al. (1989) and Roy (unpublished data).
47
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3.2 Geochemical Characteristics of
Deep-Well-Injection Zones
This section provides information on the range of en-
vironmental conditions that occur in deep-well-injection
zones in different geologic regions of the United States.
Section 3.2.1 (Lithology) discusses the types of sedimen-
tary formations that are suitable for deep-well injection
and confining layers and provides some information on
geologic formations that are currently used, or have been
used in the past, for deep-well injection of wastes. Section
3.2.2 (Brine Chemistry) discusses the typical range of
chemical characteristics of formation waters found in
injection zones.
3.2.1 Lithology
Rock that can be mapped over a large area based on
mineralogy, fossil content, or other recognizable charac-
teristic is called a formation. The lithology (texture
and mineralogy) of a geologic formation influences its
suitability for deep-well injection. Sedimentary carbonates
and sandstones usually have suitable geologic and
engineering characteristics for disposal of hazardous
wastes by deep-well injection. These characteristics in-
clude sufficient porosity, permeability, thickness, and ex-
tent to permit use as a liquid-storage reservoir at safe
injection pressures (Warner et al., 1986). In 1981, 62%
of the injection wells in the United States were drilled into
two types of reservoir rocks, either consolidated
sandstone or unconsolidated sands that had not yet
been altered by cementation to form a strongly cohesive
sandstone (see Table 3-7). The latter were usually of
Tertiary age. At that time (1981), 34% of all wells used
limestones and dolomites as reservoir rock and 4% used
miscellaneous formations. The physical and chemical
characteristics of the mineral components of sedimen-
tary formations used for injection are discussed in Sec-
tion 3.1.4 (Reservoir Matrix).
Sedimentary-rock formations that overlie the injec-
tion formation are called confining layers. To
prevent injected wastes from migrating to higher
strata or to potential underground sources of drinking
water, a confining layer must have certain geologic
and engineering characteristics:
• Sufficient thickness and area to prevent upward
migration of wastes.
• Low porosity and permeability and the ability to
maintain bw porosities and permeabilities when in-
teracting with wastes that may dissolve minerals
through neutralization.
Table 3-7 Lithology and Age of Geologic Formations
Used for Injection of Industrial Wastes
Percentage of Wells3
1967
1973 1981
Lithology
Sand
Sandstone
Subtotal
Limestone and dolomite
Other
Evaporites
Shale
Schist and gneiss
Subtotal
30
45
34
28
75
22
62
34
Age
Subtotal
Paleozoic
Permian-Mississippian
Devonian-Silurian
Ordovician-Cambrian
Subtotal
Precambrian
12
15
15
29
Total Wells
277
277
62
34
3
1
Quaternary
Tertiary
Mesozoic
Cretaceous
Triassic
o
— 27
— —
—
39
7
15
15
23
— 59
269
Percentage based on a total of 277 wells in 1967 and
1973, and 269 wells in 1981.
Sources: Adapted from Warner and Orcutt (1973) and
Reeder(1981).
Lack of natural continuous fracturing or faulting,
and resistance to artificial fracturing in response
to injection pressures.
No abandoned unplugged or improperly plugged
wells.
48
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Sedimentary rocks that are most likely to meet the
first three criteria are unfractured shale, clay,
siltstone, anhydrite, gypsum, and salt formations.
Massive limestones and dolomites (i.e., carbonates
with no continuous fracturing and solution channels)
can also serve as confining layers. Their suitability
must be determined case by case (Warner et al.,
1986). The fourth criterion has no relationship to
lithology.
Formations from all geologic periods have been used
for deep-well injection, but Paleozoic rocks are used
for most injection zones (53% in 1981), followed by
Tertiary-age formations (39%) (see Table 3-7). Older
Paleozoic rocks have been more frequently used for
injection primarily because they tend to be more
deeply buried. However, the more recent Tertiary-
age Gulf Coast sediments are also very thick, and
most injection in rocks of this age takes places there.
Dozens of specific formations have been used over
the years. Some, such as the Cambrian-age Mt.
Simon sandstone, have been used since at least the
1960s for injection of hazardous wastes in several
states (Illinois, Indiana, Michigan, and Ohio). Other
formations (such as at Wilmington, North Carolina-
see the case study in Section 7.5) have proved to be
unsuitable and have been abandoned. Table 3-8 lists
some of the formations that have been used in the
last several decades; however, not all are still used.
For example, Pennsylvania currently does not have
any Class I wells, but at least four formations were
used in the past.
Figure 3-1 shows major sedimentary basins and other
geologic features in the United States that are
significant in deep-waste injection well-site evaluations,
and Figure 3-2 provides a general indication of site
suitability based on geologic factors. A large number
of national, regional, and state waste-injection
suitability studies have been published. These are
listed in Table 3-9.
3.2.2 Brine Chemistry
Brines are classified according to their chemical con-
stituents. At least nine distinct types are recognized by
petroleum geologists (Donaldson, 1972), but most brines
encountered in injection operations are either Na-CI or
Na-Ca-CI brines (Kreitler, 1986). None is similar to
seawater, and the geochemical mechanisms by which
such brines develop are not well-understood. Three
mechanisms have been proposed to explain the high
concentrations of dissolved solids and the chemical com-
position of brines (Hanor, 1983), but at present there is no
consensus on their relative importance in explaining brine
chemistry (Kreitler, 1986). The dominant mechanism at
work in a deep-well environment has important im-
plications for the hydrodynamic conditions affecting
the movement of injected wastes. The mechanisms
and their implications are summarized in Table 3-10.
The salinity, pH, and chemical composition of the
very saline and briny waters into which hazardous
wastes are injected can vary greatly, both among
geologic basins and within a single formation.
Table 3-11 summarizes salinity and pH of some
major geologic basins and formations.
Figure 3-1 shows the locations of the basins in
Table 3-11. Maximum salinities in the Tertiary sec-
tion of the Gulf of Mexico basin (the most extensively
used strata for deep-well injection) reach almost four
times that of seawater. The Michigan basin has the
highest salinity, reaching 400,000 mg/L TDS, more
than 11 times that of seawater. In Florida, however,
where seawater circulates through the Floridan
aquifer, maximum salinities tend to be controlled by
the salinity of the seawater (Henry and Kahout,
1972).
The Frio formation, in Texas, receives more hazard-
ous waste by volume through deep-well injection
than any other geologic formation in the United
States. Table 3-11 shows that its average salinity is
about twice that of seawater (72,185 mg/L TDS), but
individual samples range from a low of 10,528 mg/L
TDS (barely above the salinity cutoff for potential
USDWs), to a high of more than 118,000 mg/L
TDS. Data from sites in Illinois and North Carolina
indicate the presence of very saline water (around
20,000 mg/L TDS, but still less saline than seawater).
The importance of pH in influencing geochemical
processes was discussed in Section 3.1.1. Table 3-11
shows that the pH of formation waters in the Frio for-
mation varies widely from moderately acidic (5.7) to
moderately alkaline (8.2), with nearly neutral averages
(6.8). The pH of formation waters from other injection
sites tends to be more alkaline, ranging from slightly
alkaline (Belle Glade, Florida, pH 7.5) and moderately
alkaline (Wilmington, North Carolina, pH 8.6), to very
alkaline (Marshall, Illinois, pH 7.1 to 10.7).
3.3 Influence of Environmental Factors
on Waste/Reservoir Compatibility
This section focuses on environmental conditions that
may result in physical or chemical incompatibilities oe-
tween wastes and reservoirs. Determining the potential
for incompatibility is a part of the geochemical fata as-
sessment that must be undertaken for any injection
project because of possible operational problems that
49
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Table 3-8 U.S.
Type of
Formation/Name
Unconsolidated
Sand
Catahoula
Cockfield
Frio
Glorieta
Miocene
Nactoch
Wilcox
Woodbine
Yegua
Sandstone
Bethel
Burgoon
Eau Claire
Eutaw
Glorieta
Granite Wash
(chert)
Greta
Mount Simon
Oriskany
Potsdam
Red Mountain
Simpson
Sylvania
Tar Springs
Theresa
Tuscarora
Yeso
Geologic Formations Being Used
Ageb State
TX
LA
TX
TX
TX
LA
Eocene AL
TX
TX
IN
PA
IN
Cretaceous FL
Permian KS
TX
TX
Cambrian IL
IL
IN
OH
PA
Ml
PA
Cambrian NY
Silurian AL
TX
Ml
IN
Cambrian NY
Silurian PA
NM
for Hazardous Waste Disposal3
# Wells0 Depth (ft)
— 3,700-5,000
— 3,000
— 5,800-7,500
— 1,300
— 3,000-6,000
— 1 ,000
2 3,400-3,800
— 2,500-5,000
— 3,400-4,500
— 2,800
— 1 ,000
— 4,000
— 3,500
2 1,100
— 5,300-5,500
— 4,500
— 4,000
2 2,600-3,100
— 5,500
6 2,800
1 —
— —
— 5,500
— 1,000-12,600
1 4,400
— 6,000-6,200
— 1,000
— 2,300
— —
— 3,700
— 1 ,000
Sources
Donaldson, 1972; Kent, 1981
Donaldson, 1972
Donaldson, 1972; Kent, 1981
Donaldson, 1972
Kent, 1981
Donaldson, 1972
Hanby et al., 1973; Hanby, 1986
Kent, 1981
ibid.
Donaldson, 1972
ibid.
ibid.
ibid.
Latta, 1973
Donaldson, 1972; Kent, 1981
Donaldson, 1972
ibid.
Brower et al., 1989
Donaldson, 1972
Clifford, 1973; Bentley et al., 1986
Reeder, 1981
Briggs, Jr., 1968
Donaldson, 1972
McCannet al., 1968
Hanby etal., 1973
Kent, 1981
Donaldson, 1972
ibid.
McCannet al., 1968
Hardaway, 1 968
Donaldson, 1972
50
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Table 3-8 (continued)
Type of
Formation/Name Age
State # Wells"
Depth (ft)
Sources
Carbonate
Arbuckle
dolomite
Bass Islands
Cedar Valley
Dundee
Ellenberger
dolomite
Eminence-Potosi
Hunton limestone
Lake City
Salem Mississippian
St. Peter
San Andres
Devonian
limestone
Virginian
limestone
Cambro-
Ordovician
Cambrian
Cambrian
Eocene
Devonian
Pennsylvanian
KS
OK
KS
PA
IL
Ml
TX
IL
TX
FL
IL
KY
TX
IL
KS
KS
— 2,000 Donaldson, 1972
— 4,000 ibid.
25 3,300-6,300 Latta, 1973
1 1,600 Donaldson, 1972; Reeder, 1981
— 2,500 Donaldson, 1972
— 4,000 ibid.
— 6,000-6,200 Kent, 1981
4 3,600-5,000 Broweretal., 1989
— 5,700-5,800 Kent, 1981
— 1,800 Donaldson, 1972
1,500-2,100 Broweretal., 1989
— 1,000 Donaldson, 1972
— 5,000 ibid.
1 2,400 Broweretal., 1989
— 3,200 Donaldson, 1972
2 3,900-4,400 Latta, 1973
Other
Wellington
(salt)
Permian
KS
220-420
Latta, 1973
aSection 3 of U.S. EPA (1985) contains a detailed compilation of injection and confining-zone characteristics (facility name,
lithology, thickness, formation name) for Class I hazardous waste injection wells as of 1983.
bAge indicated only if specified in reference source.
cDash indicates that reference did not specify number of wells.
51
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Table 3-9 Sources of National, Regional, and State Information on Suitability of Geologic Formations for
Deep-Well Injection
Geographic Area References
National
United States
Anadarko Basin
(CO, KS, OK, TX)
Appalachia
Atlantic and Gulf
Coastal Plain
Midwest
Williston Basin
(MT, ND, SD)
Love and Hoover (1960), AAPG (1968), Interstate Oil Compact Commission (1968), Piper
(1969), Rima et al. (1971), Reeder et al. (1977), Warner and Lehr (1977)
Regional
MacLachlan(1964)
Colton(1961)
LeGrand(1962)
ORSANCO (1973,1976)
Sandberg(1962)
State
Alabama
California
Colorado
Florida3
Illinois
Kansas
Michigan
Montana
New Mexico
North Dakota
New York
Ohio
Oklahoma
Oregon
Pennsylvania
South Dakota
Texas
Wyoming
Alverson (1970), Tucker and Kidd (1973), Hanby et al. (1973), Hanby (1986)
Repenning (1960)
MacLachlan (1964), Gabarini and Veal (1968), Peterson et al. (1968), Irwin and Morton (1969)
Garcia-Bengochea and Vernon (1970), Henry and Kahout (1972), Miller (1979), Vecchioli (1981)
Bergstrom (1968a,b), Bond (1972), Broweret al. (1989)
MacLachlan (1964), Edmund and Gobel (1968), Irwin and Morton (1970), Latta (1973)
de Witt (1960), Briggs (1968)
Beikman (1962), Sandberg (1962)
Repenning (1959), Peterson et al. (1968), Irwin and Morton (1970)
Sandberg (1962)
Kreidler (1968), McCann et al. (1968), Waller et al. (1978)
Clifford (1973,1975), Bentley et al. (1986)
MacLachlan (1964), Irwin and Morton (1969)
Newton (1970)
Hardaway (1968), Rudd (1972)
Sandberg (1962)
MacLachlan (1964), Irwin and Morton (1970), Kent (1981), Bassett and Bentley (1983),
Kreitler (1986), Kreitler et al. (1988)
Beikman (1962)
aSee also references for case studies in Chapter Seven, Sections 7.2, 7.3, and 7.4.
52
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Figure 3-1 Geologic Features Significant in Deep-Waste Injection-Well Site Evaluation, and
Locations of Industrial-Waste Injection Systems, 1966 (Warner, 1968).
LEGEND
1 EXTENSIVE AREAS WHERE
\ RELATIVELY IMPERMEABLE KJNEOUS-
S INTRUSIVE AND METAMOBPH1C
i EXPOSED AT SUKfACE
- BOUWAWES or GEOLOGIC FEATURES
|^' APPROXMATE BASW OUTLME5
mOUTTTHL W«3TT NJCCTKM SYSTEMS (F
WOLOCIC DETAIL NOT SHOWN
Table 3-10 Implications of Brine-Formation Mechanisms on Movement of Injected Wastes
Mechanism
Brine Type
Implications
Residual left after
precipitation of evaporites
(salt deposits).
Solution of halite present as
bedded or domal salt-evapor-
ite deposits.
Reverse osmosis. Basinal
waters forced through low-
permeability shales, leaving
the high pressure side.
Na-Ca-CI
Na-CI
Na-Ca-CI
Na-CI
Na-Ca-CI
Brines are as old as the
formation in which they occur;
stagnant conditions exist.
Active hydrologic conditions
exist, although neither the
mechanism nor the rate of fluid
movement is indicated.
Active hydrologic conditions
exist because large volumes of
water would have to pass brine
through a basin to reach
observed brine concentrations.
53
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may result from waste/reservoir incompatibility. The
major operational problems that can occur are:
• Well plugging (Section 3.3.1)
• Casing/confining layer failure (Section 3.3.2)
• Well blowout (Section 3.3.3)
In extreme situations, incompatibility between injec-
tion fluids and reservoir components can be so great
that deep-well disposal will not be the most cost-
effective approach to waste disposal. In other situa-
tions, such remedial measures as pretreatment or
controlling fluid concentrations or temperatures can
permit injection even when incompatibilities exist. In
addition to operational problems, waste-reservoir in-
compatibility can cause wastes to migrate out of the
injection zone (casing/confining-layer failure) and
even cause surface-water contamination (well
blowout).
Four major types of chemical interactions are impor-
tant when evaluating compatibility:
• Waste interactions with brine
• Waste interactions with rock
• Waste-brine mixture interactions with rock
• Microbiological interactions with the waste/brine/
rock system
Each interaction involves numerous chemical processes.
The dominance of a specific interaction depends on the
type of waste, the characteristics of the brine and rock
in the reservoir, and environmental conditions. Table 3-12
describes some of the more common processes that
may result in incompatibility.
Figure 3-2 Site Suitability for Deep-Well Injection of Industrial Waste, and Locations of
Industrial Waste Disposal Wells, 1976 (Reeder et al., 1977).
LEGEND
Unfavorable under
all conditions
Generally unfavorable but may have
limited use under restricted conditions
Favorable
controlled
under
conditions
Disposal Wells
Abandoned or plugged
Disposal Wells
54
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Table 3-1 1 Selected Parameters of
Geologic Formation and/or
Location
Major Geologic Basins3
Gulf of Mexico (Tertiary)
East Texas
Palo Duro (Texas)
Illinois
Michigan
Frio Formation (Texas)
Average (32 samples)
Minimum
Maximum
Floridan Aquifer
Maximum
Brines from Formations
Salinity (mg/L)
130,000
260,000
250,000
200,000
400,000
72,185
10,598
118,802
35,000
Used for
PH
—
6.8
5.7
8.2
Deep-Well Injection of Hazardous Wastes
Sources
Krietler, 1986
Kreitleretal., 1988
Inferred from Henry and
Belle Glade, Florida
Average (6 samples) —
Wilmington, North Carolina13
Average (4 samples) 19,150
Illinois
Marshall0 22,000
HCI injection formation 23,500
Indiana
Steel mill waste-injection formation 113,825
7.5
8.6
7.1
to 10.1
Kahout, 1972
Kaufman et al., 1973
Leenheerand Malcolm, 1973
Royetal., 1989
Kamath and Salazar, 1986
Hartman, 1968
aAII salinity figures for basins are maximums.
bSandstone, gravel, limestone aquifers.
°Devonian limestone.
Table 3-12 Processes Significant in Different Types of Waste-Reservoir Interactions
Interaction
Process
Waste with in situ fluids
Waste with rock
Waste/Brine with rock
Microbiota
Precipitation may result from incompatible brine. Hydrolysis may detoxify wastes.
Complexation may increase or decrease mobility depending on conditions. Oxidation or
reduction of wastes may occur.
Dissolution by highly acidic or alkaline wastes may threaten well and rock integrity. Case";
generated by dissolution of carbonates may cause immiscible phase separation and well
blowout. Adsorption on mineral surfaces may immobilize wastes. Clays may be mobilized
and clog pores.
Waste/brine precipitates may clog pores. Successive adsorption/desorption reactions may
occur at a particular location as waste/brine mixtures of varying proportions come in contac\
with the rock.
May form mats that clog pores'near the injection well. May transform waste to nontoxir or
other toxic forms.
55
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3.3.1 Well Plugging
The term well plugging refers to any of a variety of
processes that reduce the permeability of the injec-
tion formation or the screens that are placed in the
well's injection interval. When permeability is
reduced, injection rates must be reduced and/or
injection pressures increased. Table 3-13 lists a
number of ways in which plugging may occur. One or
more of these situations will probably take place in
most injection wells; the number and severity of
reactions will determine whether serious operational
problems arise. If plugging is confined to the immedi-
ate vicinity of the injection well, wastes will not
migrate into the injection zone until permeability is
reestablished by physical or chemical means (see
Table 3-13). Partial reductions in permeability may
allow wastes to move into the injection zone but at
increased pressures. This latter situation may con-
tribute to well-casing or confining-layer failure (Sec-
tion 3.3.2). Clay swelling, mobilization of fine par-
ticles by dissolution, and precipitation are the com-
mon causes of well plugging.
3.3.1.1 Clay Swelling
Water-sensitive clays are those which tend to swell
in the presence of water. Such swelling is most likely
to occur when wastes with low salt concentrations
replace brines in the injection zone. The reduction in
salinity causes adsorbed cations to be released from
exchange sites on the clay until concentrations in
solution and on adsorption sites are in equilibrium
again. The empty exchange sites are then hydrated,
resulting in swelling.
When clays swell, pore space is reduced, with con-
sequent reductions in permeability. Permeability
reductions will be severe when such water-sensitive
clays as montmorillonite are present, but they have
Table 3-13 Causes of Well Plugging and Possible Remedial Actions
Cause
Possible Action
Particulate solids and/or colloids.
Bacterial growth on well screen and formation.
Emulsification of two fluid phases.
Precipitates resulting from mixing of injection and
reservoir fluids.
Expansion and dispersion of water-sensitive clays
(particularly montmorillonite).
Migration of fines (very small particles) released
by dissolution.
Reprecipitation of dissolved material (iron or
calcium sulfate).
Change in wettability or reduction in pore dimensions
by adsorption (organics with large molecular weight).
Flow of unconsolidated sands into bore.
Scaling on injection equipment by precipitation from
injection fluid.
Entrapped gases.
Filter before injection.
Treat with bactericides.
Do not exceed solubility limits of organic wastes in water.
Use pretreatment or buffer of non-reactive water.
Avoid injection of low-salinity solutions in water-sensitive
formations. Use clay stabilizers.
Neutralize before injection.
Use pretreatment.
Difficult to remedy.
Use gravel-pack well screen. Inject a slug of brine after every
period of interrupted flow.
Use pretreatment; flush with solutions to remove accumulated
scale.
Remove gases from waste before injection or treat to prevent
gas formation in the injection zone.
BameS °972)l Donaldson and Johansen (1973), Hower et al. (1972), Davis and Funk (1972), Veley
(1969), and Orlob and Radhakrishna (1958).
56
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also been observed in unconsolidated sand reser-
voirs consisting of less-sensitive clays such as illite
and kaolinite (Baptist and Sweeney, 1955).
The sensitivity of clays to changes in salinity can be
drastically reduced by adding various metal and
alkaline-earth compounds, which form complex metal
ions by hydrolysis (Veley, 1969). Apparently, these
complexes are adsorbed so strongly to the clay sur-
faces that they remain adsorbed even when salinities
are reduced, thus preventing hydration of exchange
sites in the expandable lattice structure. Table 3-14
lists the effectiveness of various compounds in
preventing permeability reductions caused by clay
swelling, as shown by laboratory tests using Berea
sandstone. Some of these compounds are available
as commercial preparations to treat formations by
reducing water sensitivity. Such treatment, however, will
probably drastically reduce adsorption as a means of im-
mobilizing injected wastes, since the compounds are
resistant to ion exchange.
Additional discussion of water-sensitive clays and
responses to pH and salinity changes can be found in
the following references: Jones (1964), Land and
Baptist (1965), and Mungan (1965).
3.3.1.2 Fine-Panicle Mobilization
Highly alkaline wastes can reduce the permeability of
the injection zone by dissolving silica and releasing
fines (clay particles) that migrate and plug pores
(Hower et al., 1972). Sandstones tend to be most sen-
sitive to this reaction.
3.3.1.3 Precipitation
Many reactions between injected wastes and reser-
voir fluids can cause precipitation in deep wells
(Headlee, 1950; Selm and Hulse, 1960; Warner,
1966; Warner and Doty, 1967). Alkaline earth metals
(calcium, barium, strontium, and magnesium) can
precipitate as insoluble carbonates, sulfates,
orthophosphates, fluorides, and hydroxides. Other
metals such as iron, aluminum, cadmium, zinc,
manganese, and chromium can precipitate as
insoluble carbonates, bicarbonates, hydroxides,
orthophosphates, and sulfides. Finally, some oxidation-
reduction reaction products such as hydrogen sulfide
can precipitate.
Table 3-14 Effectiveness of Various Metal Ions in Controlling Formation-Water Sensitivity3
Reduce Sensitivity
No Effect
Increase Sensitivity
Thorium(IV) nitrate Magnesiumb
Zirconium oxychloride Barium
Aluminum6 Strontium6
Tinb
Lead"
lndiumb
Iron"
Titanium6
Hafnium6
Sodium6
Potassium6
Ammonium6
Cesium6
Lithium6
Rubidium6
aSandstone cores were treated with a sequence of fluids: (1) 3% calcium chloride brine, (2) distilled water (k0), (3) test
solution, (4) distilled water, (5) 3% sodium chloride brine, and (6) distilled water (kd). The ratio of kd/ko measured the
solution's effectiveness in preventing permeability reduction. Perfect protection would produce kd/ko - 1.0 when
cores were not stabilized with the test solution (step 3). In the base case (no test solution), kd/k0 averaged 0.003.
Reduced sensitivity: kd/ko > 0.8
No Effect: Did not cause damage or increased sensitivity to distilled water, but did not prevent damage compared to
base case.
Increased Sensitivity: Caused severe permeability damage either during injection of the test solution or during
injection of distilled water immediately after the test solution.
bSpecific compounds used were not identified.
Source: Adapted from Veley (1969).
57
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Ferric hydroxide, which is gelatinous, is more likely to
clog pores and reduce permeability than barium sulfate
and calcium sulfate, which are finely crystalline
(Warner, 1966). Above pH 10, calcium, barium, stron-
tium, magnesium, and iron all form gelatinous
hydroxide precipitates. Lower-pH solutions containing
bicarbonates will convert to carbonates if the pH is
raised, resulting in iron, calcium, and magnesium
carbonate precipitates (Strycker and Collins, 1987).
Dissolution reactions involving silica in low-pH solu-
tions can sometimes even reduce permeability by
reprecipitation elsewhere, thus clogging pores (Grubbs
et al., 1972). Table 3-15 summarizes examples of field
Table 3-15 Examples of Waste/Reservoir Incompatibility
Waste
Reservoir
Interaction
Sources
Ferric chloride
Acidic iron-
rich pickling
liquor
Acidic aluminum
nitrate radio-
active waste
Dolomite
(fractured)
Dolomitic
sandstone
Carbonates
Limonite
Ferric-hydroxide precipitate.
Neutralization with HCI was
not successful; neutralization
with acetic and citric acid
prevented precipitation but
was too expensive. Intermittent
initial injection to allow a
coating of precipitate on
fracture surfaces prevented
further precipitation.
Permeability of sandstone
unchanged, but permeability
of dolomite reduced by 1
to 3 orders of magnitude by
precipitation of iron carbonate
and/or hydroxide.
Precipitation of aluminum and
ferric oxide gels.
Hower et al.,
1972
Ragone et al.,
1978
Roedder, 1959
Phenolic wastes
(petrochemical)
Sulf uric acid
pickling liquor
Zeolite water
softener back-
wash wastes
Ferric chloride
and ammonium
hydroxide
Miocene
sands
Mt. Simon
sandstone
Arbuckle
formation
(Kansas)
Packed sand
column
Potential loss of permeability
due to clay swelling minimized
by adding 1.5% brine to waste
stream.
Precipitation of calcium
sulfate. Injection of fresh-
water buffer zone solved the
problem.
Chromates and phosphates added
to the waste water as corrosion
inhibitors formed precipitates
when they came in contact with
BaSO4, HzS, and soluble iron in
the formation waters.
Ferric hydroxide precipitate
reduced permeability of the
sand by 30%.
Sadow, 1963
Hartman, 1968
Latta, 1973
Warner, 1966
58
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and laboratory studies of waste/reservoir com-
patibility reactions, including precipitation.
3.3.2 Well-Casing and Confinlng-Formation
Failure
Interactions between corrosive wastes and casing
and packing can threaten the integrity of a well if
proper materials have not been used in construction.
Of equal concern is the potential for failure of the
confining zone due to physical or chemical effects.
For example, dissolution of an overlying carbonate
confining layer may allow upward migration of wastes.
This process was observed when hot acidic wastes
were injected in a Florida well (see case study in
Section 7.4).
Chemically active injected fluids can also have nega-
tive impacts on the mechanical properties of the reser-
voir rock. For example, adsorption of aluminum and
iron hydroxides and ferric chloride on quartz and other
silicates can weaken the surface silicon-oxygen bonds
by hydrolysis, reducing the surface energy, surface
cohesion, and breaking strength of the formation
(Swolf, 1972). In addition, stress changes caused by
increased injection pressures can fracture rock, forming
permeability channels in a confining formation through
which injected fluids could escape (Swolf, 1972).
3.3.3 Well Blowout
Gases entrapped in pore spaces resulting from phase
separation of gases from liquids (see Section 2.2.4)
can reduce the permeability of a formation (Orlob
and Radhakrishna, 1958). This process was the
major cause of clogging at ground-water recharge
wells in the Grand Prairie Region in Arkansas
(Sniegocki, 1963). Normally, pressures in deep-well-
injection zones are high enough to keep gases in
solution, so phase separation is not a problem.
However, it is possible for permeability to be reduced
by air entrainment at the same time gases are
generated by reactions between the injected waste
and reservoir formation. The resulting pressure then
forces waste and reservoir fluid up the injection well
to the surface, causing a well blowout.
The hazard of well blowout is greatest if hydrochloric
acid wastes exceeding certain temperature and con-
centration limits are injected into a carbonate forma-
tion. When carbonate dissolves in acid, carbon
dioxide is formed. Normally, this gas remains dis-
solved in the formation waters at deep-well tempera-
tures and pressures, but if temperature exceeds
88°F or acid concentration exceeds 6% HCI, carbon
dioxide will separate from the formation waters as a
gas (see Section 7.6). The resulting gas accumula-
tion can increase pressures to a point where if injec-
tion stops or drops below the subsurface carbon
dioxide pressure, a blowout can occur. Section 7.6
describes a well blowout and discusses its cir-
cumstances.
3.4 Influence of the Deep-Well
Environment on Biodegradation
Biodegradation of hazardous organic compounds in
ground waters has been the subject of much re-
search in recent years. McNabb and Dunlap, (1975)
and Ghiorse and Wilson (1988) provide good general
reviews of the topic. Unpolluted near-surface
aquifers typically contain enough oxygen for aerobic
processes to prevail. For example, Ghiorse and Wilson
(1988) summarize biodegradation data on 38 trace
organic contaminants in subsurface materials from
pristine sites. At most sites aerobic degradation is
observed. In contrast, the deep-well-injection en-
vironment typically is anaerobic (see Section 3.1.2).
This section discusses:
• Occurrence of microbes in the deep-well environ-
ment (Section 3.4.1)
• Degradation of organic compounds in anaerobic
conditions (Section 3.4.2)
• Microbial ecology of the deep-well environment
(Section 3.4.3)
3.4.1 Occurrence of Microbes
Messiniva (1962) classifies subsurface sediments and
rocks into geochemically active and geochemically in-
active categories, based on microbial activity. Geo-
chemically active sediments and rocks tend to be
heterogeneous, containing organic material, nitrogen,
and phosphorus, and support indigenous bacteria
populations. Geochemically inactive formations do not
maintain in situ microbial populations and lack fermen-
tive properties when microorganisms are added. Such
rocks typically are homogeneous, well-sorted clays
(Messiniva, 1962). As noted in Section 3.1.4.4, Sinclair
and Ghiorse (1987) describe similar relationships be-
tween microbiological activity and the saturated zone in
near-surface aquifers: gravelly sand was the most
biologically active and clayey layers the least.
It is now generally accepted that microorganisms
are ubiquitous in the deep subsurface, although, as
noted, not all strata are biologically active. Microor-
ganisms have adapted to the complete range of en-
vironmental conditions that exist on and below the
earth's surface. They have been observed at pressures
up to 25,000 psi, temperatures up to 100°C, and salt
59
-------
concentrations up to 300,000 mg/L (Kuznetsov et al.
1963). Pokrovskii (1962), using the 100°C isotherm
as the bwer boundary of the biosphere, identifies three
geothermal provinces in the European U.S.S.R.:
(1) low temperature (the Ukrainian shield), where the
isotherm lies at 10,000 to 15,000 m; (2) moderate
temperature (Russian platform), where the isotherm
lies at depths from 2,900 to 5,500 m; and (3) high
temperature (Black Sea Basin), where the isotherm
lies at 1,500 to 2,500 m, with some locations where it
reaches the surface. The currently accepted upper-
temperature limit for life is about 120°C (Ghiorse and
Wilson, 1988), which would place the limit even
deeper than that estimated by Pokrovskii, although
the diversity and activity of microorganisms at ex-
treme temperatures will surely be limited. The
temperature values reported for injection zones in
the United States (see Table 3-5) generally range
between 40° and 75°C, which is about the optimum
range for growth of thermophilic bacteria (see Sec-
tion 3.4.3).
Most pre-1970 research on microorganisms in the
deep-surface was done by petroleum microbiologists.
Dunlap and McNabb (1973) summarize data from 30
studies reporting isolation of microorganisms from
deep-subsurface sediments. Because deep-well injec-
tion zones in the Gulf Coast region (where most deep-
well injection of hazardous wastes occurs) are
commonly associated with petroleum-producing strata,
this research probably has some relevance. Sulfate-
reducing organisms are ubiquitous (Postgate, 1965).
Kuznetsov et al. (1963), in an analysis of 50 samples of
oilfield waters in the Soviet Union, found methanogenic
organisms in 23 samples. Sazonova (1962), in a study
of 18 oil deposits in the Kuibyshev region, U.S.S.R.,
found sulfate-reducing and methane-forming bacteria
to be most common; denitrifying bacteria were also fre-
quently observed. Denitrifying and sulfur-oxidizing bac-
teria are widespread in deep artesian waters in the
Soviet Union, occurring at depths exceeding 1,800 m
(Gurevich, 1962).
Ghiorse and Wilson (1988) review 14 studies, publish-
ed between 1977 and 1987, characterizing subsurface
microorganisms in pristine aquifers; only three studies
involve samples deeper than 1,000 ft below the sur-
face. Olson et al. (1981) found sulfate-reducing and
methanogenic bacteria in stratal waters from wells
1,800 m deep in the Madison Limestone in Montana.
White et al. (1983) in a comparison of microbial activity
in the Bucatanna clay at 410 m near Pensacola,
Florida, with that in the shallow Fort Polk aquifer, in
Louisiana, found the biomass to be about half that in
the shallow aquifer and found greater evidence of the
byproducts of anaerobic bacterial activity. Weirich and
Schweisfurth (1983) found 103 viable organisms/g at
a coal layer 405 m below the surface but no viable
counts in unsaturated sandstone overlying the coal, at
23 to 343 m.
Other relatively recent studies of microbial activity in
deep aquifers include DiTommaso and Elkan (1973),
Willis et al. (1975), Ehrlich et al. (1979), and Christofi
et al. (1985). DiTommaso and Elkan (1973) conducted
microbiological studies of a formation in Wilmington,
North Carolina, into which industrial waste was injected
at a depth of 850-1,000 ft (see Section 7.5). Samples
from an unpolluted observation well yielded mostly
denitrifying bacteria. No sulfate-reducers or meth-
anogens were identified in the pristine samples, al-
though significant quantities of methane production
began after the injected wastes reached the observa-
tion well (see Section 3.4.2). The authors speculated
that methanogens in the pristine samples were not
identified because they were adapted to pressures in
the formation (200 psi) and would not grow in the cul-
tures at atmospheric pressure.
Willis et al. (1975) examined samples from (1) an un-
contaminated 1,500-ft monitoring well associated
with a waste-injection facility at Pensacola, Florida
(see Section 7.2), (2) an uncontaminated 1,000-ft
monitoring well associated with the facility in North
Carolina discussed in the previous paragraph, and
(3) a 1,000-ft well near Calabash, North Carolina.
The Florida sample contained no detectable sulfate,
about 1 mg/L hydrogen sulfide, and a small amount
of dissolved nitrogen and methane. Methanogens
and denitrifiers were identified in the sample, but no
sulfate reducers. The two North Carolina wells con-
tained sulfate-reducers as well as methanogens and
denitrifiers. Inorganic constituents were not analyzed
for the well that was sampled, but four observation
wells in the vicinity of the North Carolina observation
well contained sulfate concentrations between 210
and 740 mg/L (Leenheerand Malcolm, 1973).
Ehrlich et al. (1979) examined microbial populations in
samples of industrial wastes containing acrylonitrile
and inorganic sodium salts (nitrate, sulfate, and
thiocyanate) that had been injected to a depth of 375-
425 m at second waste-injection facility at Pensacola,
Florida (see Section 7.3). Samples were obtained by
allowing the injected waste to backflow, with a maxi-
mum estimated aquifer residence time of 107 hours.
Denitrifying bacteria dominated in the waste/formation-
water mixture (105 to >106 organisms/mL), although
substantial populations of both aerobes and anaerobes
were also present (103 to 106 organisms/mL).
60
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Christofi et al. (1985, as reported by Strycker and
Collins, 1987) sampled underground mines in Europe
ranging from 600 to more than 3,000 ft deep for
microbiological activity. Salinity ranged from less
than 570 mg/L to more than 132,000 mg/L chloride.
Microorganisms were found in all samples, although
the greatest variety appeared in the least-saline
aquifer.
3.4.2 Degradation of Organic Compounds In
Anaerobic Conditions
Anaerobic biodegradation of xenobiotic (manmade) or-
ganic compounds has received less study than aerobic
biodegradation. For example, the most comprehensive
study of biodegradation of organic priority-pollutant
compounds (Tabak et al. 1981) uses aerobic condi-
tions. Kobayashi and Rittmann (1982) list about 90 ex-
amples of hazardous anthropogenic compounds and
the microorganisms that can degrade them, but less
than a third of the examples involve anaerobic
degradation.
The three most significant groups of bacteria that
may mineralize hazardous organic compounds are
(1) denitrifiers (which reduce nitrate to nitrogen),
(2) sulfate reducers (which reduce sulfate to hydrogen
sulfide), and (3) methanogens (which reduce carbon
dioxide to methane). Environmental conditions favor-
ing or restricting activity of these major groups are dis-
cussed in Section 3.4.3. Tables 3-16, 3-17, and 3-18
list published biodegradation studies under denitrifying,
sulfate-reducing and methanogenic conditions. Where
the studies report the amount of degradation at
specified time periods, these data are also included in
the tables.
Biodegradation of organic compounds under denitrify-
ing conditions (see Table 3-16) has been the least-
studied of the three groups. Ehrlich et al. (1979)
inferred that acrylonitrile injected into a carbonate
aquifer was completely degraded because the waste
was not found in samples taken from a monitoring
well where the waste arrived about 260 days after
injection began, nor in any subsequent samples (see
Section 7.3). Bouwer and McCarty (1983a) observed
partial to almost complete degradatbn of carbon
tetrachloride (> 95%), bromodichloromethane (> 55%),
dibromochloromethane (> 85%), and bromoform (> 90%)
in laboratory batch experiments simulating denitrifying
conditions. Compounds studied that did not show
significant degradation under these conditions include
chlorinated benzenes, ethylbenzene, naphthalene, chloro-
form, 1,1,1-trichloroethane, and 1,2-dibromomethane.
Phthalic acids (Aftring et al., 1981), phenol (Ehrlich et
al., 1983), tri-sodium nitrilotriacetate (Ward, 1985),
and o- and m-xylene (Kuhn et al., 1985) are other com-
pounds for which degradation has been observed
under denitrifying conditions.
Degradation of organic compounds by sulfate-
reducing bacteria (see Table 3-17) has been studied
mostly in the context of petroleum deposits (Novell!
and ZoBell, 1944; Rosenfeld, 1947; Davis, 1967).
Zajic (1969) states that these microbes are good
scavengers of organic waste products regardless of
source of waste. Novell! and ZoBell (1944) reported
finding some strains of sulfate-reducing bacteria that
use hydrocarbons, beginning with decane and
higher forms, paraffin oil and paraffin wax. In
this study, the aromatic hydrocarbons—benzene,
xylene, anthracene, and naphthalene—are not
degraded, nor are aliphatic hydrocarbons, hydrocar-
bons with molecular weight lower than that of
decane, or hydrocarbons of the naphthene series
(cyclohexane). Rosenfeld (1947) reported that high-
molecular-weight aliphatic hydrocarbons are quick-
ly decomposed by sulfate-reducing bacteria.
However, current thinking is that molecular oxygen
is required to degrade saturated hydrocarbons and
that the experiments in the above-cited papers did not
fully simulate anoxic conditions (Schink, 1988).
Only a few recent studies investigate degradation of or-
ganic compounds under sulfate-reducing condi'ions.
Gibson and Suflita (1986) found phenol to be aliiost
completely degraded under sulfate-reducing conditk ns
in three months and found that various chlorophenois
showed some degradation (not complete) during tho
same period. Smolensk! and Suflita (1987) found t'.at
sulfate-reducing bacteria degraded cresols more re idi-
ly than methanogenic bacteria, with p-cresol degraded
most readily, m-cresol less readily, and o-creso1 per-
sisting over 90 days.
Degradation of organic compounds by methanogens
has been the most extensively studied of the three
groups (see Table 3-18). Methanogenic bacteria can
readily degrade a number of monocyclic aromatics
(phenol and some chlorophenols [Gibson and Suflita,
1986] benzene, ethyl benzene [Wilson et al., 1987]
and a number of Ci and Cz halogenated aliphatic
compounds [Bouwer et al., 1981; Roberts et al.
1982; Bouwer and McCarty, 1983a; Wilson et al.,
1986]). However, the amount of degradation
depends on the specific compound and conditions
favorable for bacteria that can adapt to degrade the
compound. For example, Godsy et al. (1983) studied
biodegradation of 13 chlorophenols in the field and
laboratory, but only two (2-methylphenol and 3-
methylphenol) biodegraded significantly.
61
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Table 3-16 Organic Compounds Degraded Under Denitrifying Conditions
Compound
Aery lonit rile
Benzoate
Carbon tetrachloride
Bromodichloromethane
Dibromochloromethane
Bromoform
% De-
graded"
100
—
>95
>55
>85
>90
Time
Period8 Reference
<9mo Ehrlichetal., 1979
— Evans, 1977
3 wk Bouwer and McCarty, 1 983b
6wk
6wk
6wk
Phthalic acids
Phenol
Tri-sodium nitrilotriacetate
o-, m-Xylene
p-Cresol
Aftringetal. 1981
Ehrlichetal., 1983
Ward, 1985
Kuhnetal., 1985
Bossert and Young, 1986
aDashes indicate degradation was observed, but not quantified.
Table 3-17 Organic Compounds Degraded Under Su If ate-Reducing Conditions
Compound
Benzoate
Phenol
2-Chlorophenol
4-Chlorophenol
2,4-Dichlorophenol
2,5-Dichtorophenol
3,4-Dichtorophenol
2,4,5-Trichlorophenol
% De-
graded3
100
99
20
26
39
48
29
52
Time
Period3 Reference
3 mo Gibson and Suflita, 1986
3 mo
3 mo
3 mo
3 mo
3 mo
3 mo
3 mo
o-,m-,p-Cresols
Grease
Keratins
Organic sludge
Smolensk! and Suflita, 1987
Pipes, 1960
"Dashes indicate degradation was observed, but not quantified.
62
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Table 3-18 Organic Compounds Degraded under Methanogenic Conditions
Compound
% De-
graded3
Time
Period8
Reference
Benzoate —
m-Cresol —
o-, m-, p-Cresol —
Benzene 99
Ethylbenzene 99
o-, m-, p-Xylene —
Benzoate 100
3-Chlorobenzoate 100
3,4-Dichlorobenzoate 96
3,5-Dichlorobenzoate 100
Phenol 100
2-Chlorophenol 100
3-Chlorophenol 100
4-Chlorophenol 100
2,4-Dichlorphenol 100
2,5-Dichlorphenol 83
2,4,5-Trichlorophenol 39
Phenol —
Phenol —
2-Methylphenol —
3-Methylphenol —
Phenol —
Chlorophenols
Halogenated aliphatics —
Ci and Ca Halogenated aliphatics0
Chloroform 99
Carbon tetrachloride >99
1,1,1-Trichloroethane 97
1,1,2,2-Tetrachloroethane 97
Tetrachloroethylene 76
Bromoform >99
Bromodichloromethane >99
Dibromochloromethane >99
1,2-Dibromomethane >99
1,1-Dichloroethylene —
cis-trans-1,2-Dichloroethylene —
120wk
120wk
3 mo
3 mo
3 mo
3 mo
3 mo
3 mo
3 mo
3 mo
3 mo
3 mo
3 mo
48 hr
48 hr
48 hr
48 hr
48 hr
48 hr
48 hr
48 hr
48 hr
Evans, 1977
Ferry and Wolf e,1976
Smolenski and Suflita, 1987b
Goerlitz et al., 1985b
Wilson etal., 1987b
Gibson and Suflita, 1986
Ehrlich et al., 1983
Godsy etal., 1983
Suflita and Miller, 1985
Wood etal., 1985
Bouwerand McCarty, 1984
Bouwer and McCarty, 1983a
Bouweretal., 1981
Barrio-Lage et al., 1986°
63
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Table 3-18 (continued)
Compound
% De-
graded3
Time
Period3
Reference
Bromoform
Chloroform
Chlorodibromomethane
Dichlorobromomethane
Tetrachloroethylene
1,2-Dibromomethane
1,1,1-Dichloroethane
trans-1,2-Dichloroethane
Trichloroethylene
Styrene
1,1,1 -Trichloroethane
1,1,1-Trichloroethane
o-Xylene
Acetic acid
Formic acid
Methanol
99
99
87
66-99
99
16wk
40 wk
40 wk
40 wk
16 wk
Roberts et al, 1982°
Wilson etal., 1986b
Vogel and McCarty, 1987
Barker etal., 1986a
DiTommaso and Elkan, 1973
aDashes indicate degradation was observed, but not quantified.
Listed by Ghiorse and Wilson (1988) as carbon dioxide being dominant electron acceptor; degradation data, where given,
are also taken from this source.
Percentages shown are for acclimated methanogenic biofilm column as reported in Bouwer and McCarty (1984). Bouwer
et al. (1981), and Bouwer and McCarty (1983b) report data on different experiments that resulted in different percentages in
some instances.
As discussed in the next section, biodegradation in
ground-water systems may involve complex interac-
tions among many types of bacteria, including
denitrifying, sulfate-reducing, methanogenic, and
others. Whether complete mineralization occurs
depends on the compound, environmental condi-
tions at the site, and the microorganisms that are
best adapted to those conditions. For example,
Wood et al. (1985) identify a number of species of
facultative anaerobes that degrade certain
hatogenated aliphatic hydrocarbons (tetrachloroethyl-
ene, trichloroethylene, and 1,1,1-trichloroethane) but in
the process form intermediate hazardous compounds
that are not easily bbdegraded (vinyl chloride, 1,1 - and 1,2-
dichloroethane, and trans- and cis-1,2 dchloroethene—
see Chapter Two, Table 2-6).
Iron- and manganese-reducing and ammonia-
producing bacteria may also be significant in
biochemical reactions that occur in the subsurface
environment, but during the preparation of this refer-
ence guide no studies were identified that reported
on the degradation of hazardous organic compounds
by these bacteria. Iron and manganese oxides usual-
ly are broken down through microbial reduction
(Silverman and Ehrlich, 1964). Consequently, the
possibility of this process should be considered when
evaluating chemical reactions of iron and man-
ganese species in the deep-well environment. Lovley
(1987) reviews the literature on biomineralization of
organic matter with the reduction of ferric iron, and
Ehrlich (1987) reviews the literature on manganese
oxide reduction through anaerobic respiration.
Neither of these papers cite any studies using
xenobiotic compounds.
3.4.3 Microbial Ecology
Shturm (1962) reviewed the interactions among
ecological factors on microorganisms in oil deposits
and concluded that these interactions are not well
enough understood to establish any complete picture
64
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of their total effect on the activity of microorganisms.
In a more recent review of the biology of
methanogenic bacteria Zeikus (1977) similarly con-
cluded that relatively little is known about the chemi-
cal and mechanistic limitations of anaerobic
decomposition of organic matter in nature.
As noted in Section 3.1.4.3, the dissolved-organic-
carbon content of subsurface waters is sufficient
to maintain a small but diverse population of
microorganisms. Denitrifiers, sulfate-reducers, and
methanogens are likely to be present in low numbers
in most ground water unless conditions strongly
favoring one group exist. Consequently, when a
potential energy source in the form of an organic
contaminant enters the water, the group most
capable of utilizing the substrate at the environmen-
tal conditions existing in the aquifer will adapt and in-
crease in population, while the population of other
indigenous microbes will remain small or possibly be
eliminated.
Effects of Salinity Typical salinities in deep-well
injection zones range from about 20,000 to 70,000
mg/L (see Table 3-11), which is within the optimum
range (50,000-60,000 mg/L) for halophilic organisms
(Kuznetsov et al. 1963). Many nonhalophilic bacteria
can also live within this range. For example, a test of
14 microbe genera representing widely varying
groups showed that most grew in salt concentrations
of up to 60,000 mg/L (Hof, 1935, as cited by Zajic,
1969). Nitrification readily occurs at high salinities.
Rubentschik (1929, as cited by Zajic 1969), ob-
served conversion of ammonia to nitrate at con-
centrations of 150,000 mg/L NaCI, and isolated a
culture of Nitrosomonas showing optimal growth at
40,000 mg/L. However, very high concentrations
may slow denitrification. Hof (1935) found that it took
more than three times as long for the same amount
of gas to be generated from denitrification at 300,000
mg/L NaCI as at 30,000 mg/L NaCI (10 vs. 3 days).
Sulfate reduction readily occurs in solutions contain-
ing up to 200,000 mg/L NaCI, with an upper limit of
300,000 mg/L (Zajic, 1969)
Effects of Pressure In general, growth and reproduc-
tion of both aerobic and anaerobic bacteria occurring at
near-surface conditions decrease with increasing pres-
sures (ZoBell and Johnson, 1949). However, certain
barophilic (pressure-loving) bacteria have adapted to
the temperature and pressure conditions in the deep-
well environment. For example, aliphatic acids (acetate
ions) are degraded by methanogenic bacteria in oilfield
waters as long as temperatures are tower than 80°C
(Carothers and Kharaka, 1978). Additionally, ZoBell
and Johnson (1949) found that certain sulfate-reducing
bacteria isolated from oil-well brines located several
thousand feet below the surface are metabolically
more active when compressed to 400 to 600 atm
(40.5 to 60.8 MPa) than at 1 atm. On the other hand,
the pressures in deep-well waste injection formations
may be sufficiently high to kill or otherwise severely
affect the metabolic activity of microbes from surface
habitats that may be indigenous to the injected wastes
(McNabb and Dunlap, 1975).
Sulfate-Reducing Bacteria Sulfate-reducing bac-
teria are adapted to survive in a wide range of
anaerobic environmental conditions. The literature
on sulfate-reducers in the subsurface is extensive,
primarily because most early studies looked for only
this type of bacteria (McNabb and Dunlap, 1975).
Sulfate-reducers tolerate an Eh range from +600 to
-400 mV and salinities up to 300,000 mg/L NaCI and
can tolerate levels of heavy metals that would be
toxic to other organisms (Zajic, 1969). Booth and
Mercer (1963) found that concentrations of ionic cop-
per greater than 50 mg/L are toxic to several impor-
tant sulfur-reducing species.
Postgate (1959) states that an Eh of -200 mV or less
is required for initiation of growth. The extreme pH
limits for sulfate-reducing bacteria are 4.2 to 10.5,
with maximum growth observed at a pH of 7 (ZoBell,
1958). Neutral or slightly alkaline conditions are
preferred (Bass-Becking et al., 1960). Shturm (1962)
found that the optimum pH for sulfate-reducing bac-
teria from two different oil fields were 8.2 and 9.6 but
that bacteria from the former field would not grow at
pH 9.0. Baas-Becking and Kaplan (1956 as reported
by Shturm, 1962) found a pH of 6.2 to 7.9 and an Eh
ranging from -50 to -150 mV as most favorable for
the growth of sulfate-reducing bacteria in estuarine
environments.
Sulfate-reducing bacteria grow at temperatures from
0° to 100°C (ZoBell, 1958), although the optimum
temperature range for growth of thermophilic bacteria is
45° to 60°C (Shturm, 1962). The temperature for
maximum growth of sulfate-reducing bacteria in sea
water lies between 40° and 45°C (Shturm, 1962).
Sulfate-reducing bacteria from saline formation water
found in petroleum-bearing formations at a depth
greater than 1,000 to 2,000 m grows better at
pressures ranging from 400 to 1,000 atm than at at-
mospheric pressure (Shturm, 1962). Halophilic
sulfate-reducing bacteria with an optimum salinity
level above 120,000 mg/L have not been found
(Shturm, 1962).
Oxygen inhibits sulfate reduction, and phenols and
chlorophenol function as bactericides to sulfate
65
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reducers (Zajic, 1969). Bivalent cations may inhibit
sulfate-reducing bacteria, although the reason is
not understood (Kuznetsov et al., 1963). Porter
(1946) reports the following sequence of relative
inhibition of sulfate-reducing bacteria by cations:
Na+300 mV, and, although
methanogens were present, methane production did
not occur. Belyaev and Ivanov (1983) also reported
data on methane production from Devonian oil-bearing
sediments at a depth of 1500-1700 m in the Tartar
Republic. The formation waters were highly saline
(230,000 mg/L) and slightly acid (5.7-5.9 pH) with a
high content of organic carbon (290-315 mg/L), and
no methanogens were found. However, when sur-
face waters were flooded into oil-bearing beds to main-
tain pressure for oil production, the diluted brines
(8,000 to 58,000 mg/L, pH 6.8-7.0, and Eh +65 to +100
mV, organic carbon 5.0-16.1 mg/L) supported active
methanogenic populations (600-6,000 cells/L).
Interactions Among Microbial Groups. Decom-
position of organic matter in anaerobic environments
often depends on the interaction of metabolically dif-
ferent bacteria. Degradation in this situation is a multi-
step process in which complex organic compounds
are degraded to short-chain acids by facultative
bacteria and then to methane and carbon
dioxide by methanogenic bacteria. In these interac-
tions, methanogens may function as electron sinks
during organic decomposition by altering electron
flow in the direction of hydrogen production (Zeikus,
1977). The altered flow of interspecies hydrogen
transfer that occurs during coupled growth of
methanogens and nonmethanogens may result in
(1) increased substrate utilization, (2) different
proportions of reduced end products, (3) increased
growth of both organisms, and (4) displacement of
unfavorable reaction equilibria (Zeikus, 1977).
Ferry and Wolfe (1976) demonstrated the impor-
tance of intermediate microbial degradation steps in
the anaerobic degradation of benzoate. In mixed cul-
tures that ferment aromatic compounds to COa and
methane, the benzene nucleus is first reduced and
then cleaved to aliphatic acids by facultative Gram-
negative organisms, which are then converted to
suitable substrates for various methane bacteria to
complete the process (Evans, 1977). Wolin and
Miller (1987) present a more recent discussion of in-
terspecies relationships in methanogenesis.
Redox conditions favoring denitrification lie some-
where between those for aerobic and methano-
genic decomposition (Bouwer and McCarty, 1983b).
However, denitrification and methanogenesis are
not entirely mutually exclusive. Ehrlich et al.
(1983) observed evidence of both denitrifying and
methanogenic bacteria in phenol-depleted zones of a
creosite-contaminated aquifer and concluded that
the denitrifying bacteria contributed to degradation.
In this study, denitrifiers and iron reducers were the
dominant anaerobes in contaminated wells. Methane
66
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production was highest in the closest wells
downgradient from the contaminated site, indicating
the development of redox zones with methanogenic
conditions strongest where contaminant concentra-
tions were highest, changing to stronger denitrifying
conditions where contaminant concentrations were
lower.
Facultative-anaerobic heterotrophic bacteria from oil-
productive horizons in the Soviet Union degrade oil
accompanied by the elimination of gas containing
methane (14-35%), carbon dioxide (1.9-5.0%), hydrogen
(4.4-6.2%), and nitrogen (61-78.1%) (Shturm, 1962).
The generation of methane and nitrogen indicates that
both methanogenic and denitrifying bacteria were
probably active in the degradation process. Nazina et
al. (1985) report on an aerobic-anaerobic microbial
succession, including hydrocarbon oxidizing, sulfate-
reducing, fermenting, and methanogenic bacteria, in
water-flooded petroleum bearing rock of the Apsheron
Peninsula in the Soviet Union.
Studies by the U.S. Geological Survey at the Wil-
mington, North Carolina, deep-well waste-injection
facility (see Section 7.5) also provide evidence of
simultaneous degradation of organics by denitrifying
and methanogenic organisms (Leenheer and Malcolm,
1973). When the dilute waste front, containing organic
acids, formaldehyde, and methanol, reached the first
observation well, production of gases increased
dramatically. For a period of about 6 weeks, about half
the gas volume was methane and about a quarter,
nitrogen (Na). Two weeks later nitrogen had increased
to 62% and methane dropped to 33%, and after
another three weeks nitrogen had increased to 68%,
while methane had dropped to 12%. These relation-
ships indicate that the methanogens were more sensi-
tive to the increases in waste concentration as the
dilute front passed the observation well and more con-
centrated waste reached the site.
The inhibiting effects of sulfates on methane produc-
tion would seem to indicate that sulfate-reduction will
take place in preference to methanogenesis as long
as sulfates are present. However, the ecological
significance of sulfate reducer/methanogen inter-
relationships is not well-understood (Zeikus, 1977).
Cappenberg (1975) found that lactate metabolism of
sulfate-reducing bacteria in the upper sulfate-containing
sediment layers at Lake Vechten provided the main
energy source to acetate-fermenting methanogens lo-
cated lower down. Cappenberg (1975) also found that
in laboratory experiments, acetate utilization was great-
ly enhanced by the presence of sulfate-reducing
species with a methanogenic species compared with
the methanogenic species by itself using the same
substrate, but that the HteS produced by the sulfate
reducers was toxic to the methanogens. Smolensk!
and Suflita (1987) found that cresols degrade better
in sulfate-reducing conditions than in methanogenic
conditions and that sulfate additions increase the rate
of p-cresol metabolism in methanogenic incubations.
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76
-------
CHAPTER FOUR
GEOCHEMICAL CHARACTERISTICS OF HAZARDOUS WASTES
This chapter relates the chemical characteristics of
inorganic and organic hazardous wastes to the im-
portant fate-influencing geochemical processes oc-
curring in the deep-well environment. Section 4.1
discusses the differences between inorganic and or-
ganic hazardous wastes; Section 4.2 examines the
important properties of inorganic hazardous wastes
and provides detailed information on those identified
in 40 CFR Part 261 as hazardous; Section 4.3 fol-
lows a similar format to present information on or-
ganic hazardous wastes; and Section 4.4 suggests
resources and ways to obtain detailed information on
compounds of interest.
4.1 Inorganic vs. Organic Hazardous
Wastes
Hazardous wastes are broadly classified as either or-
ganic or inorganic. Carbon is the central building
block of organic wastes, whereas inorganic wastes
are compounds formed by elements other than
carbon (except for a few carbon-containing com-
pounds such as metal carbonates, metal cyanides,
carbon oxides, and metal carbides). Heavy metals
may straddle the definition: although usually as-
sociated with inorganics, they can also be incor-
porated into organic compounds. In fact, organic
forms of heavy metals, such as dimethyl mercury,
are often more toxic than inorganic compounds
formed by the same metal.
A major difference between organic and inorganic
hazardous wastes is that, with the exception of
cyanide, inorganics cannot be destroyed by being
broken down into nonhazardous component parts,
because at least one element in the compound is
toxic. Inorganic hazardous wastes containing toxic
elements can be transformed from a more to a less
toxic form, but can never be transformed to a non-
toxic form (see Section 4.2).
Toxic organic compounds (with the exception of or-
ganometallic compounds containing toxic metals),
however, may be rendered harmless in some cases
by being broken down into their inorganic com-
ponents: carbon, hydrogen, oxygen, and other non-
toxic elements. Most hazardous organic substances
must be manufactured under carefully controlled
conditions and are highly unlikely to form from the
basic elements of hydrogen, oxygen, and others
under uncontrolled deep-well environmental condi-
tions. Therefore, once these wastes have completely
broken down, their detoxification can be considered
permanent (see Section 4.3).
Another major difference between inorganic and or-
ganic compounds is the number of compounds. Inor-
ganic elements that exhibit toxic properties at levels
of environmental concern number in the dozens, and
only ten are regulated as hazardous wastes under
the UIC program (arsenic, barium, cadmium, chromium,
lead, mercury, nickel, selenium, thallium, and
cyanide). Additionally, the number of inorganic com-
pounds that any individual toxic element may form is
limited (probably fewer than 50). On the other hand,
the extreme versatility of carbon as a building block
for organic compounds means that literally millions
are possible, and the number that exhibit toxic
properties is probably on the order of thousands or
tens of thousands. At this time, however, the number
of organic compounds specifically regulated as haz-
ardous is fewer than 200.
Regardless of whether a waste is classified as or-
ganic or inorganic, it must have certain physical and
chemical properties to be suited for deep-well injec-
tion. Because water is the medium for injection, in-
jected wastes, whether organic or inorganic, will
typically be liquid and/or water-soluble or miscible,
and relatively nonvolatile. Table 1-1 in Chapter One
lists the major properties used to characterize haz-
ardous substances in emergency responses to spills
and indicates the physical state typically required for
deep-well injection.
77
-------
4.2 Chemical Properties of Inorganic
Hazardous Wastes
The only means by which inorganic wastes can be
rendered nonhazardous are dilution, isolation (as in
deep-well injection), in some cases changes in
oxidation state, and neutralization. As noted in
Chapter One (Section 1.2), acidic wastes made up
one-fifth of the injected waste volume and involved
one-third of the injection wells in 1983. Most of the
volume was from inorganic acids (hydrochloric, sul-
furic, and nitric). Acid-base characteristics and
neutralization are discussed in detail in Chapter Two
(Sections 2.2.1 and 2.3.1), so the remainder of this
section will focus on heavy metals and other hazard-
ous inorganics (selenium and cyanide).
Inorganic elements can be broadly classified as metals
and nonmetals. Most metallic elements become toxic at
some concentration. Table 4-1 lists 26 elements and
compounds that have been identified as toxic by various
sources. Nine elements (arsenic, barium, cadmium,
chromium, lead, mercury, nickel, selenium, and thallium)
and cyanide are defined as hazardous inorganics for
purposes of deep-well injection.
In aqueous geochemistry, the important distinguishing
property of metals is that in general they have a posi-
tive oxidation state (donate electrons to form cations in
solution), while nonmetals have a negative oxidation
state (receive electrons to form anions in solution). All
elements in Table 4-1 are metals except for arsenic,
boron, selenium, and tellurium. Figure 4-1 shows the
location of the elements in Table 4-1 within the periodic
table. In reality, there is no clear dividing line between
metals and nonmetals. For example, arsenic, which is
classified as a nonmetal, behaves like a metal in its
commonest valence states and is commonly listed as
such. Other nonmetals, such as selenium, behave
more like nonmetals.
Figure 4-1 also shows that the metals are divided into
light (also called alkali-earth metals) and heavy. All the
metals in Table 4-1 are heavy metals except for beryl-
lium and barium. Additionally, Figure 4-1 shows other
categories of elements that are or may be significant
chemically as dissolved species in deep-well-injection
zones: (1) alkali-earth metals (sodium, magnesium,
potassium, calcium, and strontium), (2) heavy metals
(manganese, iron, and aluminum, which may be sig-
nificant in precipitation reactions), and (3) nonmetals
(carbon, nitrogen, oxygen, silicon, phosphorus, sulfur,
chlorine, bromine, and iodine).
4.2.1 Major Processes and Environmental
Factors Affecting Ceochemical Fate of
Hazardous Inorganics
The major processes affecting geochemical fate of haz-
ardous inorganics are acid-base adsorption-desorption,
precipitation-dissolution, complexation, hydrolysis, oxida-
tion-reduction, and catalytic reactions. Table 4-2 lists the
processes that may be significant for hazardous
wastes in the deep-well environment and refers to sec-
tions in this reference guide where a detailed discus-
sion can be found. Tables elsewhere in the reference
guide that contain detailed information on inorganics
are also listed in Table 4-2. The significance of these
processes to inorganic wastes is discussed only briefly
here; additional information on individual elements is
given in Table 4-3.
Acid-base equilibrium is very important to inorganic
chemical reactions; Section 2.2.1 discusses the effects
of acid-base ionization. Adsorption-desorption (see
Section 2.2.2) and precipitation-dissolution (see Sec-
tion 2.2.3) reactions are also of major importance in
assessing the geochemical fate of deep-well-injected
inorganics. Interactions between and among metals in
solution and solids in the deep-well environment can
be grouped into four types: (1) adsorption (including
both physical adsorption and ion exchange) by clay
minerals (Veley, 1969) and silicates (Brown, 1979)
(see Section 2.2.2.2), (2) adsorption and coprecipitation
by hydrous iron and manganese oxides (Jenne, 1968;
Davis and Leckie, 1978a,b), (3) complexation by organic
substances such as fulvic and humic acids (see discus-
sion below), and (4) precipitation or co-precipitation by in-
corporation in crystalline minerals (see Section 2.2.3 and
Section 3.3.1.3 for additional information on common
precipitation reactions).
Solution complexation is of major importance for the
fate of metals in the deep-well environment (see Sec-
tion 2.3.2). Soluble metal ions in solution can be
divided into three major groups: (1) simple hydrated
metal tons (Veley, 1969), (2) metals complexed by inor-
ganic anions, and (3) organometallic complexes (Buffle
et al., 1984; Cabaniss et al., 1984; Raspor et al., 1984).
Figure 4-2 shows types of metal species and the range
of diameters of the different species in water. The
stability of complexes between metals and organic
matter is largely independent of ligand, and follows the
following general relationships (Fuller, 1977):
• Monovalent ions: Ag > Tl > Na > K > Pb > Cs
• Divalent ions: Pt > Pd > Hg > UOa > Cu > Ni >
Co > Pb > Zn > Cd > Fe > Mn > Sr > Ba
78
-------
Table 4-1 Inorganic Hazardous Wastes (Excluding Radioactive Elements)
Element/
Compound
Antimony
Arsenic
Asbestos
Barium
Beryllium
Bismuth
Boron
Cadmium
Cesium
Chromium
Copper
Cyanide
Gallium
Germanium
Indium
Lead
Mercury
Molybdenum
Nickel
Silver
Selenium
Tellurium
Thallium
Tin
Zinc
Zirconium
Non-
Metal Metal
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
EPA
UIC
X
X
X
X
X
X
X
X
X
X
EPA
Priority
X
X
X
X
X
X
X
X
X
X
X
X
X
X
California
List NRCC"
S
X X
S
S
S
S
X X
S
X X
S
S
S
X X
X X
S
X X
S
X
S
X S
S
S
aX = Separate report available, S = summary information in NRCC (1982).
Sources: U.S. EPA (1985); Callahan et al. (1979); 51 Federal Register 44715, December 11,1986 (California List);
National Research Council Canada (1978a) (As); National Research Council Canada (1979b) (Cd); National Research
Council Canada (1976) (Cr); National Research Council Canada (1978b) (Pb); National Research Council Canada (1981)
(Ni); National Research Council Canada (1979a) (Hg); National Research Council Canada (1982) (all others).
79
-------
• Trivalent ions: Fe > Ge > So In > Y > PI >
Ce>La
Hydration reactions between metal ions and water
affect mobility and adsorption but not toxicity (see
Section 2.3.3). Hydrolysis is particularly important in
the chemistry of cyanide.
Oxidation-reduction reactions may affect the mobility
of metal ions by changing the oxidation state (Gulens
et al., 1979). The environmental factors of pH and Eh
(oxidation-reduction potential) strongly affect all the
processes discussed above (see Sections 3.1.1 and
3.1.2 for more information). For example, the type
and number of molecular and ionic species of metals
change with a change in pH (see Figures 4-3, 4-4,
and 4-5). A number of metals and nonmentals (As,
Be, Cr, Cu, Fe, Ni, Se, V, Zn) are more mobile under
anaerobic conditions than aerobic conditions, all
other factors being equal (Fuller, 1977). Additionally,
the high salinity of deep-well injection zones in-
creases the complexity of the equilibrium chemistry
of heavy metals (Van Luik and Jurinak, 1979; Millero,
1984).
Fbrstner and Wittmann (1979) make the following ob-
servations about the general mobility of heavy metals
Table 4-2 Major Processes and Environmental
Factors Affecting the Geochemical Fate of
Inorganic Hazardous Wastes
Location of Additional
Information in this
Reference Guide
Process/Factor
Section
Tables
Processes
Acid-base equilibria 2.2.1
Adsorption-desorption 2.2.2 2-5,5-3
Precipitation-dissolution 2.2.3,3.4.1
Complexation 2.3.2
Hydrolysis 2.3.3
Oxidation-reduction 2.3.4
Catalysis 2.3.5
Environmental Factors
pH 3.1.1
Eh 3.1.2
Salinity 3.1.3
Mineralogy 3.1.4 3-2
Waste/Reservoir
Characterization 6.2,6.3
Table 4-3 Geochemical Properties of Listed Metals and Nonmetals
Property
Forms/Conditions
Mobility
Strong adsorption on Fe and Mn
oxides and hydrous oxides
Precipitation
Oxidation-reduction
Bioconversion
Cr is very mobile in neutral to alkaline conditions.
As is more mobile under anaerobic than aerobic conditions and in alkaline conditions.
Pb+2 is relatively immobile except in highly acidic environments.
Cd, Cr(IV), Hg. Ni, Se.
Cd + H2S —> CdS.
Cr + organic material —> insoluble (aerobic conditions) precipitates.
Cr(lll) hydroxide, carbonate, and sulfide precipitate (pH > 6); Cr(VI) does not
precipitate in these conditions.
Pb typically precipitates as Pb(OH)2, PbCOa, PbsfPO^aOH. NaCI increases solubility.
Ni carbonates, hydroxides, and sulfides are relatively insoluble; Ni oxides in acidic
solution may precipitate with neutralization.
Many selenium compounds can be reduced to produce elemental selenium when
exposed to organic matter in subsurface environment.
As(OH)s to As(CH3)3 (anaerobic); Hg (inorganic) to methyl mercury (anaerobic).
Source: Adapted from Strycker and Collins (1987).
80
-------
Figure 4-1 Periodic Chart of the Elements, Showing Position of Toxic Metals and Nonmetals
(Adapted from Lange's Handbook of Chemistry, 1967 edition).
IGHT METALS
I A II A
11
No
19
K
39102
37
Rb
55
Cs
4
Be
12
Mg
20
Co
40.08
38
Sr
56 57-7
[)„ S««
B° UMha*
89-10
&M
Adlnld
Set*
+5
VALENCE 0 — - i -
-5-144-1
ATOMIC NUMKKS
1
H
1.0080 _____
HEAVY METALS NON i
BRITTLE
IV B V> VIB VII B
40
Zr
91.22
1
d«
3
•
5
10
23 24 25
V Cr Mn
50.942 51.996 54.938
42
Mo
95.94
15 20 25
DUCTILE
vni
26- 27 28
Fe Co Ni
5S.B47 58.933 5871
30 35 40 a
III A IV A V
5 6
B C
low 10.811 12.011 "
MELTING 1 3 i 4
Al Sr
IB US 26.982 28.086 J
29 30 31 32
Cu Zn Ga Ge
63.54 65.37 69.72 72.59 7
47 48 49 50
Ag Cd In Sn
107.87 112.40 114.82 118.69
80 81 82
Hg Tl Pb
200.59 204.37 207.19
so 55 a u 70
INERT
GASES
AETALS
A VTA VHA
789
N O F
.007 15.999 18.998
15 16 17
p s a
0.974 32.064 35.453
33 34 35
As Se Br
4.922 78.96 79.909
51 52 53
Sb Te 1
21.75 127.60 126.90
83
Bi
08.98
7i 10 63 90
in ground water: (1) mobility tends to increase with in-
creasing salinity because alkali- and alkaline-earth
cations compete for adsorption sites on solids,
(2) change in redox conditions (lower Eh) can partly or
completely dissolve Fe and Mn oxides and liberate
other copre-cipitated metals, and (3) when natural or
synthetic complexing agents are added soluble metal
complexes may form.
4.2.2 Known Properties of Listed Hazardous
Inorganics
An extensive body of literature is available on the
chemistry of listed inorganic wastes although most of
it is oriented toward near-surface environments. For
example, Forstner and Wittmann (1979) present a
good overview of the aqueous geochemistry of metal
contaminants, and the various reports of the National
Research Council of Canada provide summaries of
the geochemistry of individual metals (see Table 4-1
for citations). Fuller (1977) contains over 200 cita-
tions on the movement of metals in soil, and Moore
and Ramamoorthy (1984a) devote individual chap-
ters to the chemistry of As, Cd, Cr, Cu, Pb, Hg, Ni,
and Zn in natural waters. One source that does dis-
cuss the chemistry of listed wastes in the deep-well
environment is Strycker and Collins (1987); informa-
tion on listed inorganic wastes from this source is
summarized in Table 4-3. Section 4.4 discusses how
to find detailed data on specific compounds.
4.3 Chemical Properties of Organic
Hazardous Wastes
Because carbon atoms can form strong bonds with
one another while combining with other elements,
the number of organic compounds is enormous.
More than two million such compounds have been
described and characterized, which is more than ten
81
-------
Figure 4-2 Types of Metal Species in Water (Forstner and Wittman, 1979).
Metal species
Range of diameters (pm) Examples
Free aquated ions
Complex ionic entities
Inorganic ion-pairs and complexes
Organic complexes,
chelates and compounds
Metals bound to high molecular
weight organic materials
Highly-dispersed colloids
Metals sorbed on colloids
Precipitates, mineral particles,
organic particles
Metals present in live and
dead biota
0.001
JD
2
£ V *
•3 c 2
3 r> S
I
-«-*
a
V
0.01
0.1
H2N
>r.vor
CUOH+, cucoj, Pb(co,)j-
AgSH°, CdQ*. Zn(OH)-
Me - OOCR1^, HgR,
CH, - C = O
O
\ /
Cu
O NHj
O = C -CH,
Me-humic/fulvic acid polymers
FeOOH, Mn(IV) hydrous oxides
Me.aq1*, Men(OH)v, MeCO3, etc.
on clays, FeOOH, organics
ZnSiO,, CuCO,, CdS in FeS,
PbS
Metals in algae
(Me = metal; R = alky 1)
Figure 4-3 Distribution of Molecular and
Ionic Species of Divalent
Cadmium at Different pH Values
(Hahne and Kroontje, 1973).
Figure 4-4. Distribution of Molecular and Ionic
Species of Divalent Lead at
Different pH Values (Hahne and
Kroontje, 1973.)
-10
-8 -6 -4
log [OH]
8
PH
10
\2
82
-------
Figure 4-5 Distribution of Molecular
and Ionic Species of Divalent
Mercury at Different pH Values
(Hahne and Kroontje, 1973).
times the total number of known compounds of all
other elements except hydrogen. The names and
quantity of hazardous organic compounds may be
bewildering to those without training in organic
chemistry. Further confusion can arise because a
single compound may have a popular name and
several technical names because of the flexibility in
the nomenclature conventions of organic chemistry.
Organic compounds can be broadly grouped into
hydrocarbons (compounds formed from only carbon
and hydrogen atoms) and their derivatives, in which
a hydrogen atom is replaced with another atom or
group of atoms, such as a functional group (e.g., an
atom or atom group that imparts characteristic
chemical properties to the organic molecules con-
taining it). Structurally, organic compounds can also
be classified as (1) straight-chain compounds,
(2) branched-chain compounds, and (3) cyclic com-
pounds. Another classification of organic compounds
divides these compounds between aromatics (those
with a six-membered ring structure in which single
and double carbon bonds alternate) and aliphatics
(those containing chains or nonaromatic rings of
carbon atoms).
The large number of organic compounds that
have been identified as hazardous makes it im-
possible to discuss individual compounds here. The
following sections provide a general overview of the
characteristics of seven major groups of hazardous
organics: (1) halogenated aliphatic hydrocarbons
(Section 4.3.1), (2) halogenated ethers (Section 4.3.2),
(3) monocyclic aromatics (Section 4.3.3),
(4) phthalate esters (Section 4.3.4), (5) polycyclic
aromatic hydrocarbons (Section 4.3.5), (6) nitrogenous
compounds (Section 4.3.6), and (7) pesticides (Sec-
tion 4.3.7). Appendix A lists more than 100 UIC-
regulated hazardous organic compounds in alpha-
betical order and gives their group as defined above,
the section in this chapter where the group is dis-
cussed, and the table in this chapter that sum-
marizes data on the compound. Appendix B contains
an alphabetized list of over 150 organic compounds
for which field or laboratory retardation factors/partition
coefficients and biodegradation studies have been
made, with reference citations for the different types
of studies.
Injected organic wastes are subject to a number of
geochemical processes in the deep-well environ-
ment. Table 4-4 lists the major ones that may be sig-
nificant and lists the sections in other chapters where
these processes are discussed in detail.
Three processes with the potential for the greatest
impact on the fate of organic wastes in the deep-well
environment are adsorption, hydrolysis, and biodegrada-
tion. The summary tables presented for each of the
seven groups of organic compounds contain the fol-
lowing information, where available:
• Ratings of the importance of adsorption, hydrolysis,
and biodegradation as fate processes, adapted by
Mills et al. (1985) from Callahan et al. (1979).
• Additional data on biodegradation from systematic
laboratory studies by Tabak et al. (1981) under
aerobic conditions (second column), and results of
biodegradation studies under anaerobic conditions
are summarized in Tables 3-16, 3-17, and 3-18 in
Section 3.4 (third column).
• An indication of whether the molecular-topology
model of Sabljic (1987) has been used to
calculate molecular-connectivity indices and soil
adsorption coefficients (koc—see Section 5.2.2.1)
for the compound; this information is useful when
the exact physical or chemical properties of the
compound are unknown and a structure-activity
approach is used to characterize the chemical
properties of the compound (see Section 4.4.4).
83
-------
Table 4-4 Major Processes and Environmental Factors Affecting the Geochemical Fate of Organic Hazardous
Wastes
Location of Additional Information in this Reference Guide
Process/Factor
Section
Tables
Processes
Acid-base equilibria
Adsorption-desorption
Complexation
Hydrolysis
Oxidation-reduction
Catalysis
Polymerization
Thermal degradation
Biodegradation
Environmental Factors
PH
Eh
Salinity
Mineralogy
Lithology
Temperature and Pressure
Waste/Reservoir Characterization
2.2.1
2.2.2
2.3.2
2.3.3
2.3.4
2.3.5
2.3.6
2.3.7
2.3.8
3.1.1
3.1.2
3.1.3
3.1.4
3.1.4
3.1.5
6.2, 6.3
2-4
2-5, 5-3, 5-4
2-8
2-7, 2-8
2-9,2-10,2-11
2-12,3-16,3-17,3-18
3-1
3-10,3-11
3-2
3-3, 3-4, 3-7, 3-8, 3-9
3-5, 3-6
3-15,6-1,6-2,6-3,
6-4, 6-5, 6-6
• Tables elsewhere in the reference guide where
additional information on the compound can be
found.
Most of the information in Tables 4-5 through 4-11 is
derived from studies oriented toward near-surface
fate processes and consequently should be interpreted
with caution. Ratings for adsorption and hydrolysis
should be generally applicable to the deep-well en-
vironment. In some instances a negative rating for
hydrolysis has been changed to positive for pur-
poses of deep-well injection because of the longer
time frame compared with that for near-surface fate
assessment (see Section 2.3.3).
Most studies of the biodegradation of hazardous or-
ganic compounds have been performed under aerobic
conditions; however, these conditions are most likely to
occur only near the injection point. (The first two
columns under biodegradation in the summary tables
present data drawn from such studies and hence may
have limited applicability in the deep-well environment.)
A compound will probably have to be susceptible to
anaerobic biodegradation for this process to be sig-
nificant in the deep-well environment, and studies on
anaerobic degradation of organic compounds are
reviewed in Section 3.4. Meikle (1972) describes
qualitative relationships in the biodegradation of 21
groups of organic compounds.
4.3.1 Halogenated Aliphatic Hydrocarbons
Hazardous halogenated aliphatic hydrocarbons in-
clude mostly straight-chain hydrocarbons (alkanes
containing single bonds, such as methane and
ethane, and alkenes containing one double bond be-
tween carbon atoms, such as ethene and propene)
in which one or more hydrogen atoms are replaced
by atoms of the halogen group of elements (fluorine,
chlorine, and/or bromine). Table 4-5 indicates the im-
portance of various fate processes for 26 hazardous
halogenated aliphatic hydrocarbons (note caveats in
Section 4.3 about interpreting this information).
Moore and Ramamoorthy (1986) review the behavior
of aliphatic hydrocarbons in natural waters.
As Table 4-5 shows, the importance of adsorption for
many of the compounds in this group is unknown. Ad-
sorption is rated as significant for three compounds
84
-------
Table 4-5 Geochemical Processes Affecting the Fate of Halogenated Aliphatic Hydrocarbons
Compound3
MCI
Adsorption Hydrolysis Biodegradation Koc
Tables
Chloromethane (methyl chloride)
Dichloromethane (methylene chloride)
Trichloromethane (chloroform)
Tetrachloromethane (carbon tetrachloride)
Chloroethane (ethyl chloride)
1,1 -Dichloroethane (ethylidene chloride)
1,2-Dichloroethane (ethylene dichloride)
1,1,1-Trichloroethane (methyl chloroform)
1,1,2-Trichloroethane
1,1,2,2-Tetrachloroethane
Hexachloroethane
Chloroethene (vinyl chloride)
1,1 -Dichloroethene (vinylidene chloride)
1,2-trans-Dichloroethene
Trichloroethene
Tetrachloroethene(perchloroethylene)
1,2-Dichloropropane
1,2-Dichloropropene
Hexachlorobutadiene
Hexachlorocyclopentadiene
Bromomethane (methyl bromide)
Bromodichloromethane
Dibromochloromethane
Tribromomethane (bromoform)
Dichlorodifluoromethane
Trichlorofluoromethane
?
?
?
+
9
+
+
9
+
+
?D
?D (An)
-D (An)
?(An)
?A (An)
?B
-B (An)
-C
-N (An)
?D
-A (An)
?A
?B
?A
+A
-A
-A
?D
-D
?A (An)
?N (An)
?A (An)
-N
y
y
y
y
2-7
2-6
2-6
1-3,2-6,5-5
2-6, 2-7
2-6
5-4
2-6
2-6
2-7
2-7
2-7, 5-5
5-5
5-5
Key:
Not likely to be an important process.
+ Could be an important fate process.
? Importance of process uncertain or not known.
(+) Revised rating in relationship to deep-well injection.
D Significant degradation, rapid adaptation (aerobic).
A Significant degradation, gradual adaptation (aerobic).
B Slow to moderate degradation, concomitant with significant volatilization.
C Very slow degradation, with long adaptation period required.
N Not significantly degraded under the conditions of test method (aerobic).
(An) Subject to anaerobic degradation (see Tables 3-16, 3-17, and 3-18).
y Soil adsorption coefficient (Koc) calculated by Sabljic (1987) from molecular connectivity index.
Additional data on all of these compounds can be found in Mabey et al. (1982). See description in Section 4.3 of sources
for each column. Note that different sources may give different ratings.
Sources: Callahan et al. (1979); Mills et al. (1985); Tabak et al. (1981); Sabljic (1987).
85
-------
(chloroethane, hexachlorobutadiene, and hexachloro-
cycbpentadiene). Hydrolysis may be an important pro-
cess for eight compounds in this group (chlorome-
thane, dichloromethane, chloroethene, 1,2-dichloropro-
pane, 1,2-dichloropropene, hexachlorocycbpentadiene,
bromomethane, and bromodichloromethane). Callahan
et al. (1979) rates bbdegradation as significant for
tetrachloroethene only, whereas Tabak et al. (1981)
found most compounds in the group are subject to
significant degradation under experimental aerobic con-
ditions. At least ten of the compounds are subject to
biodegradation under anaerobic conditions. Britton (1984)
discusses microbial degradation of aliphatic hydro-
carbons in more detail.
4.3.2 Halogenated Ethers
Ethers are either aliphatic (chain-structure) or aromatic
(ring-structure) hydrocarbons containing an oxygen
atom connected to two carbon atoms by single bonds.
In halogenated ethers, one or more halogens (chlorine
or bromine) replace hydrogen in the aliphatic or
aromatic portion of the molecule. Table 4-6 indicates
the importance of various processes for seven hazard-
ous halogenated ethers (note caveats in Section 4.3
about interpreting this information). This group contains
mostly aliphatic ethers except for 4-chlorophenyl
phenyl ether and 4-bromophenyl phenyl ether, which
are aromatic hydrocarbons.
Adsorption is very likely to be a more significant
process for the aromatic halogenated ethers than for
the aliphatic halogenated ethers. Hydrolysis is impor-
tant for two of the aliphatic ethers: bis(chloromethyl)
ether and 2-chloroethyl vinyl ether. The group appears
generally resistant to biodegradation, although under
certain conditions several may be degraded. No ex-
amples of anaerobic biodegradation of these com-
pounds were found during the literature review done to
prepare this reference guide.
4.3.3 Monocyclic Aromatic Hydrocarbons and
Halides
As mentioned, aromatic hydrocarbons have a six-
membered ring structure in which single and double
carbon bonds alternate. This ring structure tends to
be stable, so chemical reactions tend to result in the
substitution of hydrogen atoms for another atom or
functional group. Table 4-7 indicates the importance
of various fate processes for 23 hazardous monocyclic
aromatics (note caveats in Section 4.3 about inter-
preting this information). Five of these compounds
are hydrocarbons (benzene, ethylbenzene, toluene,
Table 4-6 Geochemical Processes Affecting the Fate of Halogenated Ethers
Compound9
Adsorption Hydrolysis Biodegradation
Tables
bis(Chloromethyl) ether
bis(2-Chtoroethyl) ether
bis(2-Chloroisopropyl) ether
2-Chloroethyl vinyl ether
4-Chlorophenyl phenyl ether +
4-Bromophenyl phenyl ether +
bis(2-Chloroethoxy) methane
Key:
++ Predominant fate-determinina process.
++ ?
-D
-D
+ ?D
?N
?N
? ?N
2-7
2-7
2-7
D
N
Not likely to be an important process.
Could be an important fate process.
Importance of process uncertain or not known.
Revised rating in relation to deep-well injection.
Significant degradation, rapid adaptation (aerobic).
Not significantly degraded under the conditions of test method (aerobic).
Additional data on all these compounds can be found in Mabev et al. (1982). See description in Section 4.3 of sources
for each column. Note that different sources may give different ratings.
Sources: Callahan et al. (1979); Mills et al. (1985); Tabak et al. (1981).
86
-------
Table 4-7 Geochemical Processes Affecting the Fate of Monocyclic Aromatic Hydrocarbons and Halides
Compound8
Adsorp- Hydro-
tion lysis
Biodegra- MCI
dation Koc
Tables
Benzene +
Chlorobenzene +
1 ,2-Dichlorobenzene (o-dichlorobenzene) +
1,3-DichIorobenzene(m-dichlorobenzene) + ?
1 ,4-Dichlorobenzene (p-dichlorobenzene) +
1 ,2,4-Trichlorobenzene +
Hexachlorobenzene +
Ethylbenzene ?
Nitrobenzene +
Toluene +
2,4-Dinitrotoluene +
2,6-Dinitrotoluene + ?
Phenol
2-Chlorophenol
2,4-Dichlorophenol
2,4,6-Trichlorophenol ?
Pentachlorophenol + -(+)
2-Nitrophenol
4-Nitrophenol +
2,4-Dinitrophenol +
2,4-Dimethyl phenol (2,4-xylenol)
p-Chloro-m-cresol
4,6-Dinitro-o-cresol + ?
Key:
++ Predominant fate-determining process.
Not likely to be an important process.
+ Could be an important fate process.
? Importance of process uncertain or not known.
(+) Revised rating in relation to deep-well injection.
D Significant degradation, rapid adaptation (aerobic).
A Significant degradation, gradual adaptation (aerobic).
T Significant degradation with gradual adaptation, followed
D(An)
D/A
T
?T
-T
-T
-N
?D/A(An)
-D
?D
-T
-T
+D(An)
?D(An)
++D
?D
+A
-D
-D
-D
?D
?D
-N
by toxicity.
y
y
y
y
y
y
y
y
y
y
y
y
2-8
2-6, 5-5
5-5
2-6
2-6, 5-5
2-6, 5-5
5-5
2-6, 5-5
1-3,2-4,2-8,2-11,3-7,5-4
2-4
2-4
1-3,2-4
2-4
2-4
2-4
2-4
2-4
N Not significantly degraded under the conditions of test method (aerobic).
(An) Subject of anaerobic degradation (see Tables 3-1 6, 3-1 7,
and 3-1 8).
y Soil-adsorption coefficient (Koc) calculated by Sablijic (1987) from molecular connectivity
index.
"Additional data on all of these compounds can be found in Mabey et al. (1982). See description in Section 4.3 of sources
for each column. Note that different sources may give different ratings.
Sources: Callahan et al. (1979); Mills et al. (1985); Tabak et al. (1981); Sabljic (1987).
87
-------
phenol, and 2,4-dimethyl phenol) and the rest are
halogenated or nitrogenated derivatives of benzene,
toluene, and phenol. Moore and Ramamoorthy (1984b)
review the behavior of monocyclic aromatics (Chapter
4) and phenols (Chapter 8) in natural waters.
Adsorption may be important for most of the com-
pounds in this group, whereas hydrolysis may not be a
significant process except for pentachlorophenol.
Callahan et al. (1979) rated biodegradation as
significant only for phenol, 2,4-dichlorophenol, and
pentachlorophenol. Tabak et al. (1981) found that sig-
nificant degradation with rapid or gradual adaptation
oc-curred for 15 of the 23 compounds. Anaerobic
degradation has been reported for five compounds
in this group (benzene, ethylbenzene, phenol, 2-
chlorophenol, and 2,4-dichlorophenol). Chapman (1972)
discusses in some detail the reaction sequence used
for the bacterial degradation of phenolic compounds; Gib-
son and Subramanian (1984) provide a general
review of microbial degradation of aromatic hydrocar-
bons; and Reineke (1984) reviews microbial degrada-
tion of halogenated aromatics.
4.3.4 Phthalate Esters
Esters contain a single oxygen atom attached to a
single carbon atom by a single bond, and a second
oxygen atom attached to the same carbon atom by a
double bond. Phthalate esters form when aliphatic
hydrocarbon groups replace the acidic hydrogen
atoms in phthalic acid (benzenedicarboxylic acid).
Table 4-8 indicates the importance of various
processes to six hazardous phthalate esters (note
caveats in Section 4.3 about interpreting this infor-
mation). All are subject to adsorption and are readily
biodegraded under aerobic conditions, but apparent-
ly not under anaerobic conditions. Ribbons et al.
(1984) review mechanisms for microbial degradation
of phthalates. Hydrolysis half-lives of four phthalate
esters (dimethyl phthalate, diethyl phthalate, di-n-
butyl phthalate, and di-n-octyl phthalate) are on the
order of thousands of days, which may be significant
in the time frame of deep-well injection. More discus-
sion of the chemical fate of phthalate esters in
aquatic environments can be found in Pierce et al.
(1980).
4.3.5 Polycyclic Aromatic Hydrocarbons
Polycyclic (also called polynuclear) aromatic hydrocar-
bons (PAHs) are composed of multiple rings connected
by shared carbon atoms (i.e., separate rings are com-
bined by sharing two carbon atoms). Table 4-9 indi-
cates the importance of various processes to the fate of
18 hazardous PAHs (note caveats in Section 4.3 about
interpreting this information). All these compounds are
pure hydrocarbons except for the two benzo-fluoran-
thenes, polychlorinated biphenyls (PCBs), and 2-
chloronaphthalene. Moore and Ramamoorthy (1984b)
Table 4-8 Geochemical Processes Affecting the Fate of Phthalate Esters
Compound3
Adsorption
Hydrolysis Biodegradation
Tables
Dimethyl phthalate
Diethyl phthalate
di-n-butyl phthalate
Di-n-octyl phthalate
bis(2-Ethylhexyl) phthalate
Butyl benzyl phthalate
+D
+D
+D
+A
+A
+D
2-7
2-7
2-7
2-7
Key:
D
A
Not likely to be an important process.
Could be an important fate process.
Importance of process uncertain or not known.
Revised rating in relation to deep-well injection.
Significant degradation, rapid adaptation (aerobic).
Significant degradation, gradual adaptation (aerobic).
Additional data on all of these compounds can be found in Mabey et al. (1982). See description in Section 4 3 of sources
for each column. Note that different sources may give different ratings.
Sources: Callahan et al. (1979); Mills et al. (1985); Tabak et al. (1981).
88
-------
review the behavior of PAHs (Chapter 5) and PCBs of PCBs. Hydrolysis is not significant for any com-
(Chapter 9) in natural waters. pounds in the group.
Adsorption and biodegradation under aerobic condi- 4.3.6 Nitrogenous Compounds
tions are significant for the entire group, but PAHs are The diverse nitrogenous-compounds group is corn-
generally resistent to anaerobic degradation. Safe posed of substances that have in common the substitu-
(1984) reviews the literature on microbial degradation tion of one or more nitrogen-containing functional
Table 4-9 Geochemical
Compound4
cenaphthene3
Acenaphthylene3
Fluorene3
Naphthalene
Anthracene
Fluoranthene3
Phenanthrene3
Benzo(a)anthracene
Benzo(b)fluoranthene3
Benzo(k)fluoranthene3
Chrysene3
Pyrene3
Benzo(ghi)perylene3
Benzo(a)pyrene
Dibenzo(a,h)anthracenea
lndeno(1 ,2,3-cd)pyrenea
Processes Affecting the Fate of Polycyclic Aromatic Hydrocarbons (PAHs)
Biodegra- MCI
Adsorption Hydrolysis datlon Koc Tables
+ - +D
+ - +D
+ - +A
+ - +D y 5-5
+ - +A y
+ - +A/N
+ - +D y
+ - +
+ - +
+ - +
+ - +A/N
+ - +D/N y
+ - +
+ - +
+ - +
+ - +
Polychlorinated biphenylsb + - +D/N3 y
2-Chloronaphthalene° - - +
Key:
Not likely to be an important process.
+ Could be an important fate process.
? Importance of process uncertain or not known.
(+) Revised rating in relation to deep-well injection.
D Significant degradation, rapid adaptation (aerobic).
A Significant degradation, gradual adaptation (aerobic).
N Not significantly degraded under the conditions of test method (aerobic).
y Soil-adsorption coefficient (Koc) calculated by Sablijic (1987) from molecular connectivity index.
3Based on information for PAHs as a group. Little or no information for specific compounds exists.
bBiodegradation is the only process known to transform PCBs under environmental conditions, and only the lighter
compounds are measurably biodegraded.
cPCB-related compound.
dAdditional data on all of these compounds can be found in Mabey et al. (1982). See description in Section 4.3 of sources
for each column. Note that different sources may give different ratings.
Sources: Callahan et al. (1979); Mills et al. (1985); Tabak et al. (1981); Sabljic (1987).
89
-------
groups for hydrogen in the structure. Amines are
derivatives of ammonia and contain a nitrogen atom
bonded to at least one carbon atom. Nitrosamines
are amines with a nitro (-NO2) functional group;
two are aliphatic (dimethylnitrosamine and di-n-propyl
nitrosamine) and one is aromatic (diphenylnttrosamine).
The two benzidines and 1,2-diphenyl hydrazine are
aromatic amines. Acrylon'rtrile contains the nitrile (-CN)
functional group. Table 4-10 indicates the importance of
various processes to the fate of seven hazardous
nitrogenous compounds (again note caveats in Section
4.3 about interpreting this information). Adsorption is a
significant process for all four of the aromatic amines;
hydrolysis is not. Compounds in the group are generally
not amenable to biodegradation. Acrylonitrile, however, is
readily mineralized by anaerobic denitrifying bacteria
(see American Cyanamid case study, Section 7.3).
4.3.7 Pesticides
By definition, any pesticide has toxic effects on or-
ganisms. Listed pesticides are those which combine
high toxicity with resistance to degradation in the en-
vironment. Moore and Ramamoorthy (1984b) review
the behavior of chlorinated pesticides in natural
waters.
Table 4-11 indicates the importance of various fate
processes for 15 hazardous pesticides (note caveats
in Section 4.3 about interpreting this information).
Most of these pesticides are chlorinated hydrocar-
bons. Adsorption can be an important process for
most. All except DDT, endosulfan, and heptachlor
resist hydrolysis, and most are also resistant to
biodegradation. Kearney and Kaufman (1972) review
conditions under which chlorinated pesticides are
biodegraded.
4.4 Locating Data on Specific Hazardous
Substances
The very large number of hazardous organic and in-
organic compounds precludes a detailed presenta-
tion of the characteristics of individual hazardous
wastes. Data on standard physical and chemical
properties of hazardous compounds are, however,
available in a number of standard sources. Com-
prehensive data on the behavior of specific substan-
ces in the environment usually are more difficult to
obtain. The first three sections below discuss avail-
able sources. Section 4.4.1 lists basic references for
data on physical and chemical properties, Section 4.4.2
Table 4-10 Geochemical Processes Affecting the Fate of Nitrogenous and Miscellaneous Compounds
Compound3
Adsorption
Hydrolysis
Biodegradation
Tables
Dimethylnitrosamine
Diphenylnitrosamine +
Di-n-propyl nitrosamine
Benzidine +
3,3'-Dichlorobenzidine ++
1,2-Diphenylhydrazine(hydrazobenzene) +
Acrylonitrile
Key:
++ Predominant fate-determining process.
.
?D/A
-N
?
-
?T
? D (An)
2-4
2-4
2-4
2-4
2-4
2-4
2-4
Not likely to be an important process.
+ Could be an important fate process.
? Importance of process uncertain or not known.
(+) Revised rating in relation to deep-well injection.
D Significant degradation, rapid adaptation (aerobic).
A Significant degradation, gradual adaptation (aerobic).
T Significant degradation with gradual adaptation, followed by toxicity.
N Not significantly degraded under the conditions of test method (aerobic).
(An) Subject to anaerobic degradation (see Table 3-16).
Additional data on all of these compounds can be found in Mabey et al. (1982). See description in Section 4.3 of sources
for each column. Note that different sources may give different ratings.
Sources: Callahan et al. (1979); Mills et al. (1985); Tabak et al. (1981).
90
-------
Table 4-11 Geochemical Processes Affecting the Fate of Pesticides
Compound8
Adsorption
Hydrolysis
Biodegradation
Tables
Acrolein
Aldrin +
Chlordane +
ODD +
DDE +
DDT + +
Dieldrin + -(+)
Endosulfan and endosulfan sulfate + +
Endrin and endrin aldehyde ?
Heptachlor + ++
Heptachlor epoxide +
Hexachlorocyclohexaneb +
Isophorone
Tetrachlorodibenzodioxin +
Toxaphene +
Key:
++ Predominant fate-determining process.
Not likely to be an important process.
+ Could be an important fate process.
? Importance of process uncertain or not known.
(+) Revised rating in relation to deep-well injection.
D Significant degradation, rapid adaptation (aerobic).
N Not significantly degraded under the conditions of test method (aerobic).
(An) May be degraded anaerobically (see Table 2-6).
+ D
? N (An) 2-6
?N
-N 2-6
-N
- N (An) 2-6, 2-7
- N 2-6, 2-7
+ N 2-7
?N
-N 2-7
?N
+ N
?D
-
+
Additional data on all of these compounds can be found in Mabey et al. (1982). See description in Section 4.3 of sources
for each column. Note that different sources may give different ratings.
blncludes lindane and alpha, beta, and delta isomers.
Sources: Callahan et al. (1979); Mills et al. (1985); Tabak et al. (1981).
identifies sources that may be turned to for up-to-
date information on the environmental fate of specific
substances, and Section 4.4.3 lists computerized
databases that may be useful for either.
In many cases, specific or experimentally measured
data are not available for the hazardous compound
or compounds of interest; Section 4.4.4 therefore
discusses using benchmark and structure-activity
concepts to evaluate geochemical fate when data
are unobtainable.
4.4.1 Basic References
The following is an annotated listing of basic references
and computerized databases concerning physical and
chemical properties of hazardous compounds.
Dangerous Properties of Industrial Materials, 6th Ed.
(1984), edited by N. Irving Sax, Van Nostrand Rien-
hold, Co., 135 W. 50th St., New York 10020.
This book is a single source of concise data on
the hazards of nearly 13,000 common industrial
and laboratory materials. The main section,
"General Information," gives synonyms, descrip-
tions, chemical formulas, and physical constants.
The Merck Index, 10th Ed. (1983), Merck and Co.,
Rahway, New Jersey 07065.
This book is a comprehensive encyclopedia of
chemicals, drugs, and biological substances with
9,856 listings. An extensive index and cross-index
make it easy to use.
91
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NIOSH/OSHA Pocket Guide to Chemical Hazards
(1985), U.S. Government Printing Office, Washington,
D.C. 20402
This pocket guide summarizes information from
the three-volume NIOSH/OSHA Occupational
Health Guidelines for Chemical Hazards. Data are
presented in tables, and the source includes
chemical names and synonyms, permissible ex-
posure limits, chemical and physical properties,
and other toxicological information.
OHMTADS: Oil and Hazardous Materials Technical
Assistance Data System.
This source is a computerized data-retrieval sys-
tem developed by EPA and accessible through
EPA regional offices; it is available as a computer
printout, manuals, or microfiches. A total of 126
possible information segments can be used to
retrieve data on more than 1,000 oil-based and
hazardous substances.
4.4.2 Sources of Information on Geochemical
Fate
Callahan et al. (1979) summarize the results of a
comprehensive literature search on the water-related
fate of 129 priority pollutants as of 1979. Mabey et al.
(1982) provide additional information on the same
compounds. Additionally, Appendix B presents sources
of information on the fate of a number of hazardous
organic compounds. More than 50 scientific and
trade journals are identified as having one or more
articles of interest. Journals that most frequently con-
tain papers relevant to deep-well geochemical fate
assessment include Environmental Science and
Technology, Geochimica et Cosmochimica Ada,
Ground Water, Water Resources Research, and
Water Research.
The following indexes and abstract series may be
useful in obtaining additional and more recent refer-
ences (drawn from Webster, 1987):
Biological Abstracts and Bblogical Abstracts/RRM.
Philadelphia, Pennsylvania. Biweekly; published since
1926.
This source presents abstracts from 9,000 primary
journals, monographs, symposia, reviews, reports,
and other sources. The data base revision,
BIOSIS PREVIEWS, is available through BRS
and DIALOG from 1969 on.
Chemical Abstracts. Chemical Abstracts Service,
American Chemical Society. Weekly; published since
1907.
The weekly issues include keyword and author in-
dexes. Annual and collective indexes include
author, general subject, chemical substance, for-
mula, and ring-system indexes. The Index Guide,
with its supplements, gives cross-references,
synonyms, and other information for using the
chemical-substance and general-substance indexes.
The computerized counterpart, CA SEARCH, can be
searched back to 1967.
Current Contents: Agriculture, Biology, and Environ-
mental Science. Institute for Scientific Information.
Weekly; published since 1970.
Each issue reproduces the table of contents of the
latest issues of more than 100 journals and the
contents of current books. A Title Work Index
facilitates locating desired articles or books. The
"Current Contents Address Directory, Science and
Technology" provides addresses of authors cur-
rently publishing in these fields. Also, tear sheets
or photocopies of most articles are available
through a document-delivery service called "The
Genuine Article."
Environment Abstracts. ElC/lntelligence. Monthly
(bimonthly May/June, Nov/Dec); published since
1971.
This journal covers published studies such as
conference papers, journal articles, and reports on
all environmental aspects. Each issue has a
review section followed by subject, industry,
source, and author indexes. This journal also lists
conferences and new books in print. Environment
Abstracts Annual and Environment Index provide
cumulative abstracts, and the Index also has a
useful directory listing government and non-
government organizations. The computerized
counterpart, ENVIROLINE, is available through
DIALOG and System Development Corp. (SDC)
and can be searched back to 1971.
Pollution Abstracts. Cambridge Scientific Abstracts.
Monthly; published since 1970.
This source presents abstracts and indexes for
2,500 publications worldwide on environmental
pollution. The abstracts section is arranged under
10 major headings, each containing full citations
and abstracts, followed by subject and author in-
dexes. A cumulative index is prepared annually.
92
-------
The computerized counterpart is available through
BRS and DIALOG, from 1970.
Index to Scientific and Technical Proceedings. In-
stitute for Scientific Information, Inc., published
monthly, cumulated annually; published since 1978.
The main entry gives the complete bibliographic
description of each proceeding with titles and
authors of individual papers. An extensive subject
index is also provided. Copies of papers may be
available though the Institute's Original Article
Text Service; a computer search service is also
available.
Government Reports Announcements & Index. Na-
tional Technical Information Service. Biweekly; pub-
lished since 1965.
NTIS is a central source for U.S. government-
sponsored research, development, and engineer-
ing reports and also for foreign technical reports.
Documents cited are available in microform and
paper. Report number, author, title, and other bib-
liographic information is followed by a brief
abstract. Keyword, author, report number, and
contact number indexes are prepared annually.
The computerized counterpart can be searched
back to 1964 and is available through BRS,
DIALOG, and SDC. NTIS also offers 28 weekly
abstract newsletters covering specific subject
areas, the most pertinent being Environmental
Pollution and Control: An Abstract Newsletter.
4.4.3 Computerized Databases
A large number of computerized databases can be
used to obtain data or references providing data on
specific hazardous substances. The review by Callahan
et al. (1979) mentioned earlier searched 15
databases: AGRICOLA, APTIC, ASFA, BIOSIS,
CHEM ABSTRACTS, COMPENDEX, DISSERTATION
ABSTRACTS, ENERGYLINE, ENVIROLINE, EPB, NTIS,
OCEANIC ABSTRACTS, POLLUTION ABSTRACTS,
SCIESEARCH, and SSIE CURRENT RESEARCH.
Many are available through government agencies.
The Directory of Online Data Bases (Cuandra/Elsevier,
New York, published quarterly since 1979) contains
a master index and information on individual
databases. The major private firms offering access to
a variety of databases are SDC (telephone: 1-800-
352-6689 in California; 1-800-421-7729 in the con-
tinental United States outside California), DIALOG
(1-800-334-2564), and BRS (1-800-245-4277). Syracuse
Research Corporation (1986) maintains several
environmental-fate databases.
4.4.4 Benchmark and Structure-Activity
Concepts
Where critical information is completely unavailable,
two approaches have been developed to evaluate
the fate of toxic chemicals in the environment: (1) the
benchmark concept and (2) the structure-activity
concept (Haque et al., 1980).
In the first approach, one or more benchmark chemi-
cals are selected from important classes of toxic
chemicals and their key environmental parameters
and physicochemical properties are measured.
These parameters are water solubility, vapor pressure,
hydrolysis, soil degradation, adsorption, volatilization,
photodegradation, and partition coefficient (Haque
et al., 1980). (Vapor pressure, volatilization, and
photodegradation are not significant in the deep-well
environment.) The behavior of a new chemical or a
known chemical for which data are unavailable can
then be predicted based on its structural similarity to
the benchmark.
The structure-activity approach assumes that cer-
tain properties and behaviors of compounds
depend on chemical structure. The search for, and
use of, quantitative structure-activity relationships
(QSARs) to predict the behavior of organic chemi-
cals is receiving increased attention in recent
years, primarily in the fields of pharmacology and
ecotoxicology. In an early application to environ-
mental fate assessment, Wolfe et al. (1978) use
structure-reactivity relationships to estimate
hydrolytic persistence of carbamate pesticides.
Nirmalakhandan and Speece (1988a) comprehen-
sively review recent developments in the use of
QSARs. In another paper, Nirmalakhandan and
Speece (1988b) predict aqueous solubility of 200
environmentally relevant organic chemicals based
on molecular structure.
Developments that may be particularly valuable for
deep-well injection are topological models for
biodegradation based on types of bonds and the
modulus of the atomic-charge difference across
the bonds (Deardon and Nicholson, 1986; as cited
by Nirmalakhandan and Speece, 1988a), and the
use of the molecular-connectivity index for predict-
ing partition coefficients (Sabljic, 1984 and 1987)
(see Section 5.2.2.1). Karickhoff (1984) provides
equations for estimating K0w with the addition of ring
fragments for aromatic hydrocarbons and for the ad-
dition of functional groups.
93
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Brown, D. W. 1979. Adsorption of Lead from Solution
on the Quartz- and Feldspar-Containing Silt Fraction
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Buffle, J., A. Tessier, and W. Haerdi. 1984. Inter-
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Davis, J. A., and J. O. Leckie. 1978b. Surface loniza-
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Deardon, J. C., and R. M. Nicholson. 1986. Pestic.
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Forstner, U., and G. T. W. Wittmann. 1979. Metal
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Fuller, W. H. 1977. Movement of Selected Metals,
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Waste Disposal Problems. EPA 600/2-77-020, NTIS
PB 266 905.
Gibson, D. T., and V. Subramanian. 1984. Microbial
Degradation of Aromatic Hydrocarbons. In Microbial
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Gulens, J., D. R. Champ, and R. E. Jackson. 1979.
Influence of Redox Environments on the Mobility of
Arsenic in Ground Water. In Chemical Modeling in
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American Chemical Society, Washington, D.C., pp.
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Haque, R., J. Falco, S. Cohen, and C. Riordan. 1980.
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Assessment and Screening of Toxic Chemicals. In
Dynamics, Exposure and Hazard Assessment of Toxic
Chemicals, R. Haque, ed. Ann Arbor Science, Ann
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Hahne, H. C. H, and W. Kroontje. 1973. Significance
of pH and Chloride Concentration on Behavior of
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Jenne, E. A. 1968. Controls on Mn, Fe, Co, Ni, Cu,
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Karickhoff, S. W. 1984. Organic Pollutant Sorption in
Aquatic Systems. J. Hydraulic Engineering 110:707-
735.
Kearney, P. C., and D. D. Kaufman. 1972. Microbial
Degradation of Some Chlorinated Pesticides. In
Degradation of Synthetic Organic Molecules in the Bio-
sphere. National Academy of Sciences, Washington,
D.C., pp. 166-188.
Mabey, W. R., et al. 1982. Aquatic Fate Process
Data for Organic Priority Pollutants. EPA 440/4-81-
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Meikle, R. W. 1972. Decomposition: Qualitative Relation-
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Millero, F. J. 1984. The Activity of Metal Ions at High
Ionic Concentrations. In Complexation of Trace
Metals in Natural Waters, C. J. M. Kramer and J. C.
Duinker, eds. Martinus Nijhoff/Dr. W. Junk Publish-
ers, The Hague, pp. 187-201.
Mills, W. B., et al. 1985. Water Quality Assessment:
A Screening Procedure for Toxic and Conventional
Pollutants (Revised 1985). EPA/600/6-85/002a-b.
Moore, J. W., and S. Ramamoorthy. 1984a. Heavy
Metal in Natural Waters: Applied Monitoring and Im-
pact Assessment. Springer-Verlag, New York.
Moore, J. W., and S. Ramamoorthy. 1984b. Organic
Chemicals in Natural Waters: Applied Monitoring and
Impact Assessment. Springer-Verlag, New York.
National Research Council Canada. 1976. Effects of
Chromium in the Canadian Environment. NRCC
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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.
National Research Council Canada. 1979a. Effects
of Mercury in the Canadian Environment. NRCC
Report No. 16739, Ottawa, Ontario.
National Research Council Canada. 1979b. Effects
of Cadmium 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
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Nirmalakhandan, N. N., and R. E. Speece. 1988a.
Structure-Activity Relationships: Quantitative Techni-
ques for Predicting the Behavior of Chemicals in the
Ecosystem. Environ. Sci. Technol. 22:606-615.
Nirmalakhandan, N. N., and R. E. Speece. 1988b.
Prediction of Aqueous Solubility of Organic Chemi-
cals Based on Molecular Structure. Environ. Sci.
Technol. 22:328-338.
Pierce, R. C., S. P. Mathur, D. T. Williams, and M. J.
Boddington. 1980. Phthalate Esters in the Aquatic
Environment. NRCC Report No. 17583, National Re-
search Council of Canada, Ottawa, Ontario.
Raspor, B., H. W. Niirnberg, P. Valenta, and M.
Branica. 1984. Significance of Dissolved Humic Sub-
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Waters, C. J. M. Kramer and J. C. Duinker, eds.
Martinus Nijhoff/Dr. W. Junk Publishers, The Hague,
pp. 317-327.
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Aromatic Compounds. In Microbial Degradation of Or-
ganic Compounds, D. T. Gibson, ed. Marcel Dekker,
Inc., New York, pp. 319-360.
Ribbons, D. W., P. Keyser, R. W. Eaton, B. N.
Anderson, D. A. Kunz, and B. F. Taylor. 1984.
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Sabljic, A. 1987. On the Prediction of Soil Sorption
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96
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CHAPTER FIVE
METHODS AND MODELS FOR PREDICTING THE GEOCHEMICAL FATE
OF DEEP-WELL-INJECTED WASTES
5.1 Basic Approaches to Geochemical
Modeling
The geochemical interactions possible between an in-
jected waste and the reservoir rock and its associated
fluids can be quite complex. Thus a combination of
computer modeling, laboratory experimentation, and
field observation will inevitably be necessary to satisfy
current regulatory requirements for a geochemical
no-migration petition. This chapter covers the computer
methods and models available for predicting geo-
chemical fate and includes the following topics:
• Basic approaches to geochemical modeling, in-
cluding:
— Model evaluation (Section 5.1.1)
— Model deficiencies (Section 5.1.2)
• Specific methods and models, including:
— Computer codes for modeling aqueous and solution
geochemistry (Section 5.2.1)
— Methods and models for predicting adsorption
(Section 5.2.2)
— Quantitative and qualitative models for predicting
biodegradation (Section 5.2.3)
— Equations for predicting hydrolysis (Section 5.2.4)
— Chemical transport modeling (Section 5.2.5)
Laboratory methods for geochemical fate assessment
are covered in Chapter Six, and field methods are
briefly discussed in Section 7.1 of Chapter Seven.
5.1.1 Model Evaluation
The American Society for Testing and Materials (ASTM,
1984) has developed a standard protocol for evaluating
environmental chemical-fate models, along with defini-
tions of basic modeling terms, shown in Table 5-1.
Predicting fate requires natural phenomena to be
described mathematically. The expression of chemical
fate can be computerized using a code (see computer-
code definition in Table 5-1) to perform the computations
and predict the results when inputs simulating conditions
of interest are provided.
Two critical aspects of the use of computer codes for
predicting geochemical fate are the verification and
validation of the models on which the codes are based.
In most cases, verifying geochemical codes (testing for
internal consistency) is relatively straightforward.
Validation studies, in which the model predictions are
compared with empirical results, however, have been
restricted mostly to simple or partial systems, and only
qualitative validation was achieved. The limited number
of validation studies raises serious questions regarding
the reliability of thermodynamic data and the current
understanding of geochemical processes occurring in
the deep-well environment (Apps, 1988).
5.1.2 Model Deficiencies
In addition to the general lack of validation, some
serious deficiencies remain (Apps et al., 1988):
• The data on thermodynamic properties of many
relevant water-miscible organic species are either
incomplete or unavailable.
• Many minerals are solid solutions (e.g., clays, am-
phiboles, and plagtoclase feldspars). Solid-solution
models either have not yet been developed or ap-
propriate algorithms have not been incorporated
into computer codes.
• Models describing the adsorption of water-miscible
organic compounds on natural materials are in the
preliminary stages of development and have not
been correlated with field observations under typi-
cal injection-zone conditions. Few computer codes
contain algorithms for calculating the distribution of
97
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Table 5-1 Definitions of Terms Used in Chemical-Fate Modeling
Term
Definition
Algorithm The numerical technique embodied in the computer code.
Calibration A test of a model with known input and output information that is used to adjust or estimate factors for
which data are not available.
Computer code The assembly of numerical techniques, bookkeeping, and control languages that represents the model
from acceptance of input data and instruction to delivery of output.
Model An assembly of concepts in the form of a mathematical equation that portrays understanding of a
natural phenomenon.
Sensitivity The degree to which the model result is affected by changes in a selected input parameter.
Validation Comparison of model results with numerical data independently derived from experiment or
observation of the environment.
Verification Examination of the numerical technique in the computer code to ascertain that it truly represents the
conceptual model and that there are no inherent numerical problems associated with obtaining a
solution.
Source: ASTM (1984).
species between the adsorbed and aqueous
states.
Calcium-sodium-chloride-type brines (which occur
typically in deep-well injection zones) require
sophisticated electrolyte models to calculate their
thermodynamic properties. Many parameters for
characterizing the partial molal properties of the
dissolved constituents in such brines have not
been determined. (Molality is a measure of the
relative number of solute and solvent particles in
a solution and is expressed as the number of
gram-molecular weights of solute in 1,000 grams
of solvent.) Precise modeling is limited to rela-
tively low salinities (where many parameters are
unnecessary) or to chemically simple systems
operating near 25°C.
Current computer codes usually calculate only
the thermodynamically most stable configuration
of a system. Modifications can simulate nonequi-
librium but there are limitations on the extent to
which codes can be manipulated to simulate
processes that are kinetically (rate) controlled:
the slow reaction rates in the deep-well environ-
ment compared with ground-water movement
(i.e., failure to attain local homogeneous or
heterogeneous reversibility within a meter or so
of the injection site) creates particular problems.
Little is known about the kinetics of dissolution,
precipitation, and oxidation-reduction reactions in the
natural environment. Consequently, simulating
the kinetics of even more complicated injection-
zone chemistry is very difficult.
Bergman and Meyer (1982) point out a particularly
relevant problem with mathematical models. The rela-
tive reliability of mathematical models (compared with
physical [microcosm] models based on empirical field
or laboratory studies) decreases rapidly as the number
of environmental pollutants being modeled increases
(see Figure 5-1). Consequently, mathematical models
tend to be less cost effective for complex waste
streams than are physical (empirical) models.
5.2 Specific Methods and Models
This section examines methods and models available
to predict the processes discussed in Chapter Two.
Most of the chemical processes discussed there (acid-
base equilibria, precipitation-dissolution, neutralization,
complexation, and oxidation-reduction) are interrelated,
i.e., reactions of one type may influence other types of
reactions, and consequently must be integrated into
aqueous- and solution-geochemistry computer codes
(see Section 5.2.1). Not all aqueous-geochemistry
codes handle adsorption; methods for predicting ad-
sorption are discussed in Section 5.2.2. Quantitative
and qualitative methods for predicting biodegradation
are discussed in Section 5.2.3. It is possible to predict
hydrolysis reactions for hazardous wastes without
complex computer codes; the necessary equations are
98
-------
Figure 5-1 Relative Trade-offs Between Physical (Microcosm) and Mathematical Models as Affected
by Effluent Complexity (Bergman and Meyer, 1982).
. , >- A
IS
Si
Ld _l
cr LJ
tr
THE ECOSYSTEM-*
PHYSICAL MODEL-
UJ >
> H
P <->
-------
• The B-dot extension of the D-H equation
• The Davies equation
• Pitzer interaction parameter equation (Pitzer and
Mayorga1973).
All three are empirical equations that can be used to
predict activity coefficients at high ionic concentrations.
The first two equations are applicable up to 0.5 molal
ionic concentrations (approximately 29,000 mg/L
NaCI), and the third can be used for extremely high
salinities (30 molal). Thermodynamic codes may be
used separately to generate needed basic ther-
modynamic data or may be incorporated as sub-
routines of aqueous geochemistry codes. Table 5-2
lists two thermodynamic codes that may be useful
when calculating thermodynamic data for geochemical
modeling (several distribution-of-species codes incor-
porate such codes). Robie et al. (1978) summarize
thermodynamic data for 133 oxides and 212 other
minerals, including properties at higher temperatures,
where available.
5.2.1.2 Distribution-of-Species Codes
Distribution-of-species codes, also called equilibrium
codes, solve a simultaneous set of equations that
describe equilibrium reactions and mass balances of
the dissolved elements. The output of these equations
is the theoretical distribution of the aqueous species for
the dissolved elements. Most codes indicate the
saturation state with respect to the solid phase, and
many also include equations to describe ion exchange
and simple linear adsorption. Two basic approaches
are used to model species distribution: equilibrium con-
stants and Gibbs free-energy minimization. The Gibbs
free-energy approach has theoretical and computation-
al advantages, but is limited by its lack of accurate
and internally consistent thermodynamic data
(Nordstrom et al., 1979). Consequently, most codes
use the equilibrium-constant approach. One such is
SOLMNEQ (see Table 5-2), which has been used by
several researchers to simulate deep-well-injection
geochemical interactions (Ehrlich et al., 1979; Roy et
al., 1989). Van Luik and Jurinak (1979) also have
used this approach, along with the cluster integral
expansion theory of electrolyte-solution structure, to
model the equilibrium chemistry of lead, cadmium,
copper, and zinc in brines (2 to 6 molal) at tempera-
tures in the 10° to 35°C range.
5.2.1.3 Reaction-Progress Codes
Reaction-progress codes, also called mass-transfer
codes, calculate both the equilibrium distribution of
aqueous species (as in distribution-of-species codes)
and new compositions of the water as selected
minerals and compounds are precipitated or dis-
solved. The more sophisticated codes incorporate
the reaction-path concept, in which incremental
steps toward equilibrium are considered along a
chosen path of mineral-water reaction. The most
versatile and best-documented of this type of code
for deep-well conditions is EQ3/6, developed at the
Lawrence Livermore National Laboratory. PHREEQE,
developed by the U.S. Geological Survey (USGS) has
recently been modified to incorporate the Pitzer interac-
tion equations (Crowe and Longstaffe, 1987; Plummer
et al., 1988). ECES is a proprietary code with similar
capabilities. See Table 5-2 for additional descriptions of
these codes.
5.2.2 Adsorption
Mineral surfaces on which adsorption may occur are
diverse and complex (see discussion of Reservoir
Matrix, Section 3.1.4.2), and the mechanisms by
which a hazardous constituent may attach to the
solid surface vary substantially (see Section 2.2.2).
Therefore, theoretical models that can be used readi-
ly to predict adsorption for a variety of compounds
over a range of conditions are difficult to develop.
Table 5-3 summarizes the applicability of three major
methods for predicting adsorption in the deep-well
environment. These methods include adsorption
isotherms, the clay ion-exchange model, and the
triple-layer model.
5.2.2.1 Adsorption Isotherms
The simplest and most widely used method for predict-
ing adsorption is to measure adsorption isotherms
(the variations in the amount of a substance adsorbed
at different concentrations measured at a constant
temperature). Empirical constants can be calculated
from such measurements. (See Section 6.4.1, for a
detailed discussion of methods.) The amount of adsorp-
tion at concentrations other than those which were
measured can then be predicted using the empirical
constants in an appropriate formula. The correct ap-
plication of this method requires acknowledging such
effects as matrix and temperature.
Three types of adsorption isotherms are discussed in
this section: (1) the linear distribution coefficient,
(2) the Langmuir adsorption isotherm, and (3) the
Freundlich adsorption isotherm. The distribution
coefficient assumes that adsorption is linear (i.e., the
amount of adsorption is directly proportional to the
concentration of the compound in solution) and is ac-
tually a special case of the Langmuir and Freundlich
isotherms, which are nonlinear (Rao and Davidson,
1980).
100
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Table 5-2 Aqueous- and Solution-Geochemistry Models of Potential Value for Modeling Deep-Well Injection
Name/Developer(s)
Description/Comments
Thermodynamic Codes
SUPCRT
Shock and Helgeson (1988,1990)
Tanger and Helgeson (1988)
PHAS20
Haas (1974)
Haas and Fisher (1976)
Can be used to calculate dissolution reaction constants at any specified
temperature between 0° and 800°C and 1 to 5,500 bars.
Developed by USGS forthermodynamic calculations.
Distribution-of-Species Codes
SOLMNEQ
Kharaka and Barnes (1973)
Handles temperatures of 0°-350°C, pressures from 1 -1,000 bars,
and salinities up to about 35,000 mg/L It includes organic complexes
and ion-exchange equilibria. The model was used by Ehrlich et al. (1979)
and Roy et al. (1989) to simulate injected waste/reservoir interactions.
Reaction-Progress Codes
EQ3/6
Walters and Wolery (1975)
Wolery(1979)
Wolery (1983)
Jackson and Wolery (1985)
Wolery (1986)
PHREEQE
Parkhurst et al. (1980)
Plummer et al. (1983)
Plummer and Parkhurst (1985)
PHRQPITZ
Plummer et al. (1988)
ECES
Scrivner et al. (1986)
Handles temperatures of 0°-300°C, and pressures of 1-500 bars.
Earlier version handles salinities up to about 0.5 molal (-29,000 mg/L);
latest version contains Pitzer interaction electrolyte model.
Has been used to model geochemical evolution of Gulf Coast
(Apps et al. 1988) and to simulate evolution of ground waters in basalt
(Solomon, 1986). Most thoroughly documented of available models.
Temperature range is 0°-100°C, at 1 bar up to about 0.5 molal.
Has successfully modeled the evolution of ground water with the
mineralogy of a limestone and dolomite aquifer in Florida.
Incorporates Pitzer interaction electrolyte model in PHREEQE for
temperatures up to about 60°C. Code has been partially validated in
laboratory studies at 52°C and an ionic strength of 5.8-9.2 molar
(personal communication, W. R. Roy, Illinois State Geological Survey,
Champaign, Illinois, May 10, 1990).
Temperature range is 0°-200°C; pressure range is 0-200 atm;
ionic strength is 0-30 molal. It incorporates the Pitzer interaction
electrolyte model for high salinities. It is a proprietary model licensed
by OLI Systems, Morristown, New Jersey.
Sources: Nordstrom et al. (1979); Apps (1988); Plummer et al. (1988).
101
-------
Table 5-3 Applicability of Methods and Models for Predicting Adsorption in the Deep-Well Environment
Method/Model
Applicability
Methods
Adsorption isotherms
Linear distribution coefficient
Langmuir
Freundlich
Relatively easy to measure. The main disadvantage is that the empirical
coefficients may change with changing environmental conditions,
requiring measurement.
Applicable only at very dilute concentrations of organic compounds and
where >0.1% organic matter is present. Usefulness is uncertain.
Underlying assumptions for the derivation of the equation typically
will not apply.
Limited available data on adsorption under simulated deep-well conditions
are best described by the formula; however, the disadvantage of all
adsorption isotherms applies.
Models
Clay ion-exchange model
Triple-layer model
May be useful for predicting adsorption of heavy metals. Aqueous-phase-
activity solid-solution model coefficients can be obtained from distribution-
of-species models. Estimating clay-phase activity coefficients is more
problematic.
Of limited value because of the complexity of adsorption sites,
unpredictable interactions among adsorbents, and complications
introduced by high salinities.
Linear Distribution Coefficient. The simplest type
of isotherm is the linear-distribution coefficient, Kd
(Apps, 1988). It is also called the partition coefficient,
Kp (Mills et al., 1985). The equation for calculating
adsorption at different concentrations is:
= KdC
[5-1]
where:
S = amount adsorbed (micrograms [u.g]/g solid)
C = concentration of adsorbed substance in
solution(u.g/milliliter [mL])
Kd = distribution coefficient
This equation is widely used to describe adsorption in
soil and near-surface aquatic environments. Another
widely used linear coefficient is the organic-carbon par-
tition coefficient (Koc), which is equal to the distribution
coefficient divided by the percentage of organic carbon
present in the system (Koc = Kd/% organic carbon), as
proposed by Hamaker and Thompson (1972). Sabljic
(1987) presents very accurate equations for predict-
ing the Koc of both polar and nonpolar organic
molecules based on molecular topology, provided the
organic matter percentage exceeds 0.1%. Karickhoff
(1984) discusses in detail adsorption processes of or-
ganic pollutants in relation to Koc to adsorption proces-
ses of organic pollutants.
Winters and Lee (1987) describe a physically based
model for adsorption kinetics for hydrophobic organic
chemicals to and from suspended sediment and soil
particles. The model requires determination of a single
effective diffusivity parameter, which is predictable from
(1) compound solution diffusivity, (2) octanol-water par-
tition coefficient, and (3) adsorbent organic content,
density, and porosity.
Major problems are associated with using the linear-
distribution coefficient for describing adsorption/
desorption reactions in ground-water systems. Some of
these problems include:
• The coefficient actually measures multiple
processes (reversible and irreversible adsorption,
102
-------
precipitation, and coprecipitation). Consequently,
it is a purely empirical number with no theoretical
basis on which to predict adsorption under differ-
ing environmental conditions or to give informa-
tion on the types of bonding mechanisms
involved.
• The waste-reservoir system undergoes a dynamic
chemical evolution in which changing environ-
mental parameters may result in variations of
Kd values by several orders of magnitude at dif-
ferent locations and at the same location at dif-
ferent times.
• All methods used to measure the Kd value in-
volve some disturbance of the solid material and
consequently do not accurately reflect in situ con-
ditions.
Apps et al. (1977) and Reardon (1981) discuss the
problems in using distribution coefficients.
Langmuir Isotherm. The Langmuir equation was
originally developed to describe adsorption of gases
on homogeneous surfaces and is commonly ex-
pressed as follows:
• Adsorption occurs only on localized sites with no in-
teractions among adjoining adsorbed molecules
• The maximum adsorption capacity (Smax) repre-
sents coverage on only a single layer of molecules
In a study of adsorption of organic herbicides by
montmorillonite, Bailey et al. (1968) found that none
of the compounds conformed to the Langmuir ad-
sorption equation. Of the 23 compounds tested only
a few did not conform well to the Freundlich equa-
tion.
Freundlich Isotherm. The assumptions mentioned
above for the Langmuir isotherm generally do not
hold true in a complex heterogeneous media such as
soil (Rao and Davidson, 1980). The deep-well en-
vironment is similarly complex and consequently the
few studies of adsorption in simulated deep-well con-
ditions (Donaldson and Johansen, 1973; Donaldson
et al., 1975; Collins and Crocker, 1988) have fol-
lowed the form of the Freundlich equation:
S = KC
N
[5-3]
= 1/kSmax + 1/CSmax
[5-2]
where:
Smax = maxmum adsorption capacity (u.g/g soil)
k
= Langmuir coefficient related to adsorption
bonding energy (mL/g)
= amount adsorbed (p.g/g solid)
= concentration of adsorbed substance in
solution
S
C
A plot of C/S versus 1/C allows the coefficients k and
Smax to be calculated. When kC « 1 , adsorption will
be linear as represented by Equation 5-1 .
The Langmuir model has been used to describe
adsorption behavior of some organic compounds
at near-surface conditions (Alben et al. 1988).
However, three important assumptions must be made:
• The energy of adsorption is the same for all sites
and is independent of degree of surface coverage
where S and C are as defined in Equation 5-1 and K
and N are empirical coefficients. Taking the logarithms
of both sides of Equation 5-3:
log S = log K + N log C
[5-4]
Thus, log-log plots of S versus C provide an easy
way to obtain the values for K (the intercept) and N
(the slope of the line). The log-log plot can be used
for graphic interpolation of adsorption at other con-
centrations, or, when values for K and N have been
obtained, the amount of adsorption can be calculated
from Equation 5-3. Figure 5-2 shows an example of
adsorption isotherms for phenol adsorbed on Frio
sandstone at two different temperatures. (Note that
when N = 1, Equation 5-3 simplifies to Equation 5-1
[i.e., adsorption is linear]). This simplified form is
used by Lindstrom et al. (1971) to model transport of
chemicals in saturated soils. However, Lindstrom et
al. (1971) state that it is the most specialized and
least generally applicable of the three mathematical
models they developed.
103
-------
Figure 5-2 Freundlich Isotherm for Phenol Adsorbed on Frio Core (Collins and Crocker, 1988).
o>
c
o
^•-*
2
.*-•
c
Q)
U
C
O
o
T)
QJ
.a
o
t/>
TJ
100
1000
Equilibrium Concentration, mg/L
10000
Weaknesses in Nonlinear Adsorption Isotherms.
The Langmuir equatbn has a strong theoretical basis,
whereas the Freundlich equation is an almost purely
empirical formulation because the coefficient N has
embedded in it a number of thermodynamic parameters
that cannot easily be measured independently (Apps,
1988). These two nonlinear isotherm equations have
most of the same problems discussed earlier in relation
to the distribution-coefficient equation. All parameters
except adsorbent concentration (C) must be held con-
stant when measuring Freundlich isotherms, and sig-
nificant changes in environmental parameters, which
would be expected at different times and locations in
the deep-well environment, are very likely to result in
large changes in the empirical constants. Alben et al.
(1988) discuss sources of uncertainty and bias in ex-
perimental Langmuir and Freundlich isotherms.
Table 5-4 shows maximum measured values reported
in three studies for adsorption of a variety of organic
compounds tested at simulated deep-well temperature
and pressure conditions. It illustrates variations that
can occur among compounds adsorbed on the same
geologic materials and variations that can occur for the
same compound adsorbed on different geologic
materials. Table 5-4 shows that the amount of adsorp-
tion at a given concentration for different organic com-
pounds varies by a factor of 24 (50 milligrams per
kilogram [mg/kg] for phenol to 1,200 mg/kg for
n-hexylamine at 10,000 mg/L in the Cottage Grove
Sandstone). Adsorption can also vary greatly for the
same compound depending on the lithology of the
sample. For example, the amount of phenol ad-
sorbed on the Cottage Grove and Frio formations dif-
fers by a factor of five at the same concentration
level (10,000 mg/L) and the amount of 1,2-
dichloroethane differs by a factor of two. Further-
more, adsorption of one compound, 1-butanol,
differed by a factor of 10 (30 versus 300 mg/kg) on
two separate experiments with the same rock forma-
tion.
An assumption implicit in most adsorption studies is
that adsorption is fully reversible. In other words, once
the empirical coefficients are measured for a particular
substance, Equations 5-1 to 5-4 describe both adsorp-
tion and desorption isotherms. This assumption is not
always true. The problem of irreversible adsorption is
discussed in detail in Section 2.2.2.3. Collins and
Crocker (1988) observed apparently irreversible ad-
sorption of phenol in flowthrough adsorption experi-
ments involving phenol interacting on a Frio
sandstone core under simulated deep-well tempera-
tures and pressures. If adsorption-desorption is not
fully reversible, it may be necessary to use separate
Freundlich adsorption- and desorption-isotherm equa-
tions to model these processes in the deep-well en-
vironment (Apps, 1988).
The most extensively studied adsorption-desorption
phenomena have been related to the adsorption of
pesticides on soils. A number of kinetic, equilibrium,
and nonequilibrium models have been developed for
pesticide-soil interactions (Van Genuchten et al.,
104
-------
Table 5-4
Results of Adsorption Experiments with Organic Compounds at Simulated Deep-Well Conditions
Lithology/
Compound
Temp.
°C
Pressure
(psi)
Concen-
tration
(mg/L)
Amount
Adsorbed
(mg/kg)
Source
Cottage Grove Sandstone
Phenol
1-Butanol
n-Hexylamine
Butanal
Cottage Grove Sandstone
Phenol
1 -Butanol
n-Hexylamine
1,2-Dichloroethane
2-Butanone
Crotonaldehyde
1-Nitropropane
Propylproponoate
Pyridine
Frio Sandstone
Phenol
1,2-Dichloroethane
66
66
66
66
60
66
66
60
60
60
60
60
60
60
38
2,940
2,940
2,940
2,940
3,000
3,000
3,000
3,000
3,000
3,000
3,000
3,000
3,000
3,400
3,400
10,000
5,000
10,000
10,000
10,000
5,000
10,000
5,000
5,000
10,000
10,000
1,500
10,000
10,000
5,000
55
300
1,200a
175
50
30
1,200 a
300
60
330
210
55
290
Donaldson and
Johansen, 1973
ibid.
ibid.
ibid.
Donaldson et
al., 1975
ibid.
ibid.
ibid.
ibid.
ibid.
ibid.
ibid.
ibid.
276 Collins and
Crocker, 1988
150 ibid.
aAdsorption rate
curve.
curves and Freundlich isotherm plots do not agree in either reference; value taken from the adsorption rate
1974; Rao et al., 1979; Rao and Davidson, 1980).
Unfortunately, little work has been done to evaluate
their applicability to the deep-well environment.
5.2.2.2 Clayton-Exchange Model
As noted above, adsorption isotherms largely are
derived empirically and give no information on the
types of adsorption that may be involved. Scrivner et
al. (1986) have developed an adsorption model for
montmorillonite clay that can predict the exchange of
binary and ternary ions in solution (two and three
ions in the chemical system). This model would be
more relevant for modeling the behavior of heavy
metals that actively participate in ion-exchange reac-
tions than for organics in which physical adsorption
is more important (see Section 2.2.2)
The clay ion-exchange model assumes that the inter-
actions of the various cations in any one clay type
can be generalized and that the amount of exchange
will be determined by the empirically determined cation-
exchange capacity (CEC) of the clays in the injection
zone (see Chapter Three, Table 3.2, for data on the
CEC of various clay types). The aqueous-phase ac-
tivity coefficients of the cations can be determined
from a distribution-of-species code (see Section 5.2.1).
The clay-phase activity coefficients are derived by
assuming that the clay phase behaves as a regular
solution (Garrels and Christ, 1965; Hildebrand et al.,
1970) and by applying conventional solution theory
to the experimental equilibrium data in the literature.
Scrivner et al. (1986) compared the ion-exchange
model predictions with several sets of empirical data.
The model predictions are very accurate for binary-
105
-------
exchange reactions involving the exchange of nickel
ions for sodium and potassium ions on illite and less
accurate for ternary reactions involving hydrogen,
sodium, and ammonia ions. The deep-well environ-
ment, however, is very likely to have multiple ex-
changeable species (such as Na+, K+, Ca+2, and
Mg+ ), and injected wastes commonly have elevated
concentrations of more than one heavy metal (see
Chapter One, Table 1-3). These concentrations result
in complex ion-exchange interactions that probably ex-
ceed the capabilities of the model.
5.2.2.3 Triple-Layer Model
One of the more sophisticated models for describing
adsorption phenomena in aqueous solutions is the
triple-layer model (TLM), also called the Stanford
General Model for Adsorption (SGMA) because it has
been developed, refined, and tested over a number of
years by faculty and researchers at Stanford University
(Davis and Leckie, 1978, 1980; Kent et al., 1988). The
TLM separates the interface between the aqueous
phase and the adsorbent surface into three layers: sur-
face layer, inner diffuse layer, and outer diffuse layer.
Each has an electrical potential, charge density,
capacitance, and dielectric constant. Hydrogen ions
are assumed to bind at the surface plane; electrolyte
ions (such as Na+) bind at the inner diffuse plane. The
surface is assumed to be coated with hydroxyl groups
(OH") with each surface site associated with a single
hydroxyl group. The hydroxyl-occupied surface sites
may either react with other ions in solution or dissociate
according to a series of reactions, with each having an
associated equilibrium constant. Experimental terms
relate the concentrations of the ions at their respective
surface planes to those in the bulk solution. The sum of
the charges of the three layers is assumed to be zero
(i.e., the triple layer is electrically neutral). For all its
sophistication, TLM currently is of limited value for
predicting adsorption in deep-well environments
(Apps, 1988):
• Site-binding constants have been determined for
only a limited range of simple oxides with only
one type of surface site. Multiple-surface site
minerals occurring in the deep-well environment
such as silicates, aluminosilicates (e.g., feldspars),
and complex oxides (such as manganese oxide)
will require much more complex TLMs. Data ade-
quate to characterize their behaviors do not exist.
• Fixed-charge minerals such as clay are even
more complex than the multiple-surf ace site
minerals, and both ion exchange and other types
of adsorption must be measured to characterize
absorption reactions fully. Again, data are not
available to predict adsorption by class over a
wide range of clay compositions and environ-
mental conditions, and the outcome of studies to
develop these data is uncertain.
• Minerals with different adsorptive properties in
the injection zone may interact to produce results
different from those which would be obtained if
each mineral were tested separately. No satisfac-
tory model has been developed that predicts ad-
sorption properties of mixtures based on the
properties of individual adsorbents.
• The TLM is based on laboratory measurements
of adsorption on materials that are suspended in
solution. No satisfactory methods for measuring
and interpreting the adsorptive properties of in-
tact host rock have been developed for TLM ap-
plication.
• The TLM has been developed using studies
based on solutions of relatively tow concentra-
tions of dissolved compounds. The very saline
and briny conditions found in the deep-well en-
vironment may require an entirely different
model.
5.2.3 Biodegradation
This section examines two quantitative models for
predicting biodegradation: the kinetic rate expres-
sions (Section 5.2.3.1) and the biofilm model (Sec-
tion 5.2.3.2). It also examines several qualitative
models for describing biodegradation in the deep-
well environment (Section 5.2.3.3).
5.2.3.1. Kinetic Rate Expressions
When microorganisms use an organic compound as
a sole carbon source, their specific growth rate is a
function of chemical concentration and can be
described by the Monod kinetic equation. This
equation includes a number of empirical constants
that depend on the characteristics of the microbes,
pH, temperature, and nutrients (Callahan et al.,
1979). Depending on the relationship between sub-
strate concentration and rate of bacterial growth, the
Monod equation can be reduced to forms in which
the rate of degradation is zero order with substrate
concentration and first order with cell concentration,
or second order with concentration and cell con-
centration (Paris et al., 1981).
The Monod equation assumes a single carbon
source. The difficulty in handling multiple carbon
sources, which are typical in nature, has led to the
use of an empirical biodegradation rate constant KB:
106
-------
ksC
[5-5]
This equation is of the same form as Equation 5-1 for
linear adsorption. Predicting biodegradation using
such a rate constant is complicated when multiple
biodegradable compounds are present. For example,
phenol and naphthalene are both rapidly biodegraded
in single-compound laboratory shake-flask experiments
when seeded with bacteria from an oil-refinery settling
pond, but when the two compounds are combined,
naphthalene is not degraded until the phenol is gone
(Bergman and Meyer, 1982).
When a compound is co-metabolized (degraded but
not used as a nutrient), a second-order biodegrada-
tion coefficient can be used to estimate ks:
KB = kB2B
[5-6]
where kB2 is an empirical coefficient and B is the bac-
terial concentration. Mills et al. (1985) describe the use
of these formulations to predict aerobic biodegradation
in surface waters and present methods of adjusting for
temperature and nutrient limitations. This approach to
predicting biodegradation is problematic because it is
difficult to obtain empirical coefficients in the deep-well
setting.
Baughman et al. (1980) derive a second-order
kinetic rate expression as a special case of the
Monod kinetic equation. It appears to describe
biodegradation of organics in natural surface waters
reasonably well:
-d[C]/dt = k[B][C]
[5-7]
Paris et al. (1981) found that degradation of several
pesticides (2,4-DBE, Malthion, and CIPC) in samples
from over forty lakes and rivers fits this second-order
model of microbial degradation.
General degradation rate models of organics in soils
have been described by Hamaker (1972), Goring et
al. (1975), Hattori and Hattori (1976), Larson (1980),
and Rao and Jessup (1982). In most instances,
biodegradation is the major, but not necessarily the
only, process affecting the rate of degradation.
5.2.3.2. Bioftlm Model
The most sophisticated model available for predict-
ing biodegradation of organic contaminants in sub-
surface systems is the biofilm model, presented by
Williamson and McCarty (1976a,b) and refined over
several years by researchers at Stanford University
and the University of Illinois/Urbana (Rittmann et al.,
1980; Rittmann and McCarty, 1980a,b; McCarty et
al., 1981; Bouwer and McCarty, 1984; Chang and
Rittmann, 1987a,b).
The biofilm model is based on two important features
of the ground-water environment: (1) nutrient con-
centrations tend to be tow, and (2) the solid matrix has
a high specific surface area. These characteristics
favor the attachment of bacteria to solid surfaces in the
form of biofilms so that nutrients flowing in the ground
water can be used (ZoBell, 1937; 1943). The presence
of low nutrient levels in the ground water also implies
that bacteria regularly must use many different com-
pounds as energy sources and, consequently, may
select organic contaminants more readily as nutrients
(Bouwer and McCarty, 1984).
The basic biofilm model (Williamson and McCarty,
1976a,b) idealizes a biofilm as a homogeneous
matrix of bacteria and the extracellular polymers that
bind the bacteria together and to the surface. A
Monod equation describes substrate use; molecular
diffusion within the biofilm is described by Pick's
second law; and mass transfer from the solution to
the biofilm surface is modeled with a solute-diffusion
layer. Six kinetic parameters (several of which can
be estimated from theoretical considerations and
others of which must be derived empirically) and the
biofilm thickness must be known to calculate the
movement of substrate into the biofilm.
Rittmann and McCarty (1980a,b) have developed
equations for incorporating bacterial growth into the
model, allowing the steady-state utilization of substrate
materials to be predicted. They also show theoretically
and verify experimentally that there is a substrate con-
centration threshold (Smin) below which no significant
activity occurs (Rittman and McCarty 1981). McCarty et
al. (1981) introduce the idea of secondary substrate
utilization by a biofilm, in which microbes can metabo-
lize trace compounds (S < Smin) in the presence of
another substrate that is in sufficient concentrations to
support biofilm growth. Bouwer and McCarty (1984) in-
corporate steady-state utilization of secondary sub-
strates into the model by coupling the bbfilm mass
(controlled by degradation of the primary substrate)
with concentration and individually determined rate
parameters for each secondary substrate. Laboratory
tests of degradation on a variety of chlorinated
107
-------
benzenes, nonchlorinated aromatics, and halogenated
aliphatics as secondary substrates agreed reasonably
well with predicted values (Bouwer and McCarty,
1984). The most recent refinement of the model incor-
porates the effects of adsorption of material substrate
to the surface on which the biofilm is attached, but is
restricted to biofilms on activated carbon (Chang and
Rittmann, 1987a,b).
When water containing substrate concentrations
greater than Smin is injected into the subsurface, the
model predicts that biofilm development will occur only
in the first meter or so of the injection zone (Rittmann et
al., 1980). Low concentrations of hazardous com-
pounds will be significantly degraded as secondary
substrates only if they are readily biodegraded in the
biofilm zone. Any amount not biodegraded in the
biofilm zone will tend to persist once it leaves the zone
of concentrated biological activity. When substrate con-
centrations are not sufficient to sustain biofilm develop-
ment, Bouwer and McCarty (1984) suggest that a
simple biodegradation coefficient such as that dis-
cussed earlier (Equation 5-5) is probably adequate.
The biofilm model has not been applied to fate as-
sessments of deep-well-injected hazardous wastes.
Its greatest potential use for modeling degradation in
the deep-well environment may be to predict the
conditions under which excessive biofilm develop-
ment might occur, with associated pore clogging. As
noted elsewhere (Table 3-13), chemical treatment of
injected fluids with biocides to reduce bacterial ac-
tivity is a common practice. At the other extreme,
highly acid or alkaline wastes can prevent bacterial
growth entirely, or concentrations of a specific com-
pound can inhibit bacterial growth through toxic ef-
fects. For example, Elkan and Horvath (1977) found
that formaldehyde concentrations in a simulated in-
jected waste significantly reduced biodegradation of
acetate.
5.2.3.3. Qualitative Models
Several qualitative models for biodegradation in the
deep-well environment have been suggested. They do
not allow quantitative predictions to be made, but they
do provide insight into the types of biodegradation
processes that may occur. These models have not
been expressed quantitatively to simulate degradation,
although relatively simple codes using first-order
biodegradation constants (ks) could probably be
developed without much difficulty. In the absence of
quantitative models for predicting biodegradation,
laboratory simulations must be used to assess
biodegradation potential (see Section 6.4.3).
The conceptual geochemical model of acidic waste after
injection into the subsurface, proposed by Leenheer and
Malcolm (1973), involves a moving front of microbial ac-
tivity (see Wilmington, North Carolina, case study, Sec-
tion 7.5) with five zones as shown in Figure 5-3: (1) a
dilute zone, controlled by diffusion, (2) a zone where
substrate concentrations are sufficiently high to allow sig-
nificant microbial activity, (3) a transition zone, where in-
creasing waste concentrations create unfavorable
conditions for microbial growth, (4) a neutralization zone,
where abiotic chemical reactions predominate, and (5) a
waste-storage zone where undiluted waste no longer
reacts with the host rock. This model implies that the rate
of injection far exceeds the zone's capacity for
biodegradation.
Figure 5-3 Proposed Geochemical Model of
Waste after Injection into the
Subsurface (Leenheer and
Malcolm, 1973).
Zones
Obs.
Well
Front
(Degradation)
Bouwer and McCarty (1984) suggest a qualitative
model that represents nonbiofilm microbial biodegrada-
tion over increasing distances from the injection point.
This model follows the redox reaction sequence dis-
cussed in Section 2.3.4. Table 5-5 shows the
progression that would occur as Eh declines with dis-
tance from the injection point and lists hazardous or-
ganic compounds that would be degraded most
readily in each zone. This model implies that most
compounds not degraded in their appropriate zone
will move through the ground-water system without
significant additional degradation. The model also
implies, however, that those compounds which are
biodegraded by methanogenesis will continue to
move through the ground water until degradation is
complete.
108
-------
Table 5-5 Redox Zones for Biodegradation of Organic Micropollutants
Increasing Distance from Injection Point •
Biological Conditions
Aerobic
heterotrophic
respiration
Chlorinated
benzenes
Ethylbenzene
Styrene
Naphthalene
Denitrification
Sulfate
Respiration
Methanogenesis
Organic Pollutants Transformed
Carbon tetrachloride
Bromodichloromethane
Dibromochloromethane
Bromoform
(see also Table 3-16)
None specified
(see Table 3-17)
Ci and C2
Halogenated
aliphatics
(see also
Table 3-18)
Source: Adapted from Bouwer and McCarty (1984).
5.2.4 Hydrolysis
Hydrolysis (see Section 2.3.3) is easily predicted,
provided that the rate constants for a compound are
known. The rate of abiotic hydrolysis is given by:
R = -kHCT
[5-8]
where:
R = the rate of hydrolysis, mole liter"1 sec"1 or
u.g liter"1 sec"1
KH = specific hydrolysis rate constant, sec"1
CT = the dissolved plus adsorbed phase
concentration of compound C, mole liter" or
u.g liter"1
The hydrolysis rate constant, kH, is actually the sum
of three rate constants:
ka = the acid-catalyzed hydrolysis rate constant,
liter mole"1 sec"1
[Hi
kb
= the concentration of hydrogen ion, mole
liter"1 ([Hi = 10"pH)
= the base-catalyzed hydrolysis rate constant,
liter mole"1 sec"1
[OH" ] = the concentration of hydroxide ion, mole
liter"1 ([OH" ]^10[pH"14])
Note that in an acid solution, kb = 0, and in an alkaline
solution, ka = 0. KH can be adjusted to include the ef-
fects of adsorption by multiplying (ka[H1 + kb[OH" ])
times the decimal fraction of the total amount of a
dissolved compound, C (Mills et al., 1985). At any fixed
pH, the half-life of a substance is independent of
concentration and can be calculated with the equation:
ti/2 = 0.693/kH
[5-10]
kH = kn + ka[H+] + kb[OKT
[5-9]
where:
kn = the neutral hydrolysis rate constant for the
pH-independent reactions of a chemical with
water, sec"1
Hydrolysis is strongly pH-dependent, with ka dominant
at low pH and kb dominant at high pH; at pH 7, kn can
often be most important. However, the detailed
relationship of pH and rate depends on the specific
values of kn, ka, and kb. If these rate constants are
known, then the hydrolysis rate at any pH can be readi-
ly calculated. Mabey and Mill (1978) provide these data
109
-------
for a large number of organic compounds, and El-
lington et al. (1986, 1987,1988) provide data on about
70 regulated hazardous pollutants.
Wolfe et al. (1978) use structure-reactivity relationships
to estimate hydrolytic persistence of carbamate pes-
ticides. Perdue and Wolfe (1983) develop a mathemati-
cal model based on application of the Bronsted
equations for general acid-base catalysis, used to
forecast the maximum contribution of buffer catalysis in
pollutant hydrolysis reactions. They conclude that at
the low concentrations of Bronsted acids and bases in
most aquatic environments, buffer catalysis is probably
insignificant.
Mills et al. (1985) describe step-by-step procedures for
calculating kH, and Scrivner et al. (1986) describe in
detail the modeling of cyanide and nitrite hydrolysis
in the deep-well environment.
5.2.5 Chemical Transport
It is beyond the scope of this reference guide to dis-
cuss chemical-transport models in detail, but basic
approaches and important models will be addressed
briefly. Currently three major approaches can be
used to modeling chemical transport:
• Retardation-factor models, which incorporate a
simple retardation factor derived from a linear- or
linearized-distribution coefficient.
• Integrated models, in which all mass, momen-
tum, and energy-transfer equations, including
those in which chemical reactions participate, are
solved simultaneously for each time step in the
evolution of the system.
• Two-step models, which first solve mass
momentum and energy balances for each time
step and then reequilibrate the chemistry using a
distribution-of-species code.
Empirically determined retardation factors (either
partition coefficients as discussed in Section 5.2.2.1
or breakthrough curve measurements, which are the
change in solute concentration measured over time
in laboratory or field experiments) have been widely
used because of their inherent simplicity (Javandal et
al. 1984). Modeling of specific geochemical partition
and transformation processes is not necessary if the
retardation factor can be determined empirically.
The problems with linear-distribution coefficients dis-
cussed in Section 5.2.2.1 apply equally to any retarda-
tion factor derived from them. Field measurements can
be made but are expensive to obtain and highly site
specific. Nevertheless, retardation factors provide some
insight into organic chemical transport. Winters and
Lee (1987), in a study of mobility of chlorobenzene,
naphthalene, and 4-chlorobiphenyl in ground-water
discharge to a sandy stream bed, found that the
measured retardation of these compounds generally
agreed with that predicted using Kows and the organic
carbon content of the sediment material. However, the
large amount of tailing observed in the organic-tracer
breakthrough curve resulted in center-of-mass retarda-
tion factors up to about two times greater than peak-to-
peak retardation factors. This discrepancy underscores
the importance of understanding the tailing phenomena
in the field before retardation factors are used for model-
ing (Winters and Lee 1987).
Integrated and two-step chemical-transport models in-
corporate distribution-of-species or reaction-progress
codes into hydrologic transport codes. The few studies
in which the two approaches have been tested using
the same set of field data have agreed reasonably well;
thus one approach does not have an obvious ad-
vantage over the other. The two-step approach tends
to be computationally less intensive than the integrated
approach but may have difficulty maintaining mass
balance when rapid precipitation and dissolution occur
(Apps, 1988).
Tables 5-6 and 5-7 present a number of models of both
types that have been described in the literature. Of the
models listed in these tables, DYNAMIX would appear
to have the greatest potential for use in simulating
chemical transport in the deep-well environment be-
cause it incorporates the reaction-progress code
PHREEQE, which can handle deep-well temperatures
(see Table 5-2). PHREEQE, however, does not incor-
porate pressure equilibria.
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110
-------
Table 5-6 Integrated Ground-Water Chemical-Transport Models
Developers Description/Comments
Rubin and James, 1973 Simulates heterovalent ion exchange and changing concentrations of
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Table 5-7 Two-Step Ground-Water Chemical-Transport Models
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111
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117
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CHAPTER SIX
FIELD SAMPLING AND LABORATORY PROCEDURES AND PROTOCOLS
6.1 Overview
Laboratory procedures for geochemical-fate assess-
ments of deep-well-injected wastes can be classified
into two categories: (1) routine chemical and physical
analyses that characterize the waste and reservoir rock
(quantities and types of different constituents), and
(2) laboratory studies that simulate the geochemical
processes in the injection zone. The first type of proce-
dures can be used to identify the types of processes
that are likely to be important. They also provide data
on basic parameters of the deep-well injection system
for hydrologic and geochemical modeling. The second
type may help predict how one or more processes will
operate in the deep-well environment and may be
necessary for verifying the results of geochemical
modeling. In the absence of geochemical models,
laboratory studies can provide empirical data on
specific geochemical processes and their interactions.
This chapter provides an overview of these procedures
with references for additional information.
6.1.1 Chapter Organization
Section 6.2 (Waste-Reservoir Characterization) lists
the physical and chemical properties that may re-
quire measurement before the wastes, reservoir
rock, and reservoir fluids can be characterized.
Section 6.3 (Waste-Reservoir Interaction Tests) ex-
amines the various methods for testing the compatibility
of wastes and injection-zone materials and for testing the
geochemical interactions among wastes, reservoir rock,
and fluids. The types of tests reported in the published
literature are also summarized in this section.
Section 6.4 (Geochemical Processes) briefly discusses
methods for measuring adsorption and rate constants
for chemical transformations, such as hydrolysis and
the biodegradability of organic wastes.
Section 6.5 (Quality Assurance/Control Procedures)
presents procedures used in sampling and laboratory
tests that ensure data reliability.
Section 6.6 (Annotated Bibliography) summarizes
references that contain detailed guidance in sam-
pling methods, analytic protocols, and interpretation
of results, and tells how to obtain these documents.
The listings in this section are also indexed by topic
(see Table 6-9).
The references included in this chapter are likely to be
useful in a wide variety of situations because (1) they
have been cited most frequently in the literature on
deep-well injection, and (2) many describe methods
designed specifically for use with hazardous wastes
and/or deep-well injection.
6.1.2 Selecting Sampling Methods and
Laboratory Procedures
Because of the variety of hazardous wastestream
compositions and the number of possible variations
in the lithology and brine chemistry of the injection
zone, no single set of sampling methods and
laboratory protocols can cover all situations. For cer-
tain parameters, more than one analytic method and
more than one standard procedure may be available.
For example, procedures recommended by the
American Public Health Association (APHA), the
American Society for Testing and Materials (ASTM),
and the U.S. Geological Survey (USGS) for measur-
ing total dissolved solids all differ slightly (Hem,
1970). Moreover, some state regulatory agencies
have their own protocols.
Given the multiplicity of laboratory procedures
and protocols, the following steps are suggested
for selecting the appropriate methods. The tables
referred to are presented in later sections and
discussed in detail.
1. Select Parameters. Identify the physical and
chemical parameters of the waste and reservoir
solids and fluids to be measured. Table 6-1 lists
basic parameters for characterizing waste-
water. Important physical properties of reser-
voirs are listed in Table 6-2, methods for
119
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analyzing reservoir rock in Table 6-3, and
chemical properties of reservoir rock and fluids
in Table 6-4. The greatest judgment will be re-
quired in selecting chemical parameters repre-
senting reservoir fluids because of the large
number of species that may be present in these
fluids. Table 6-5 provides guidance in class-
ifying dissolved species, and Table 6-6 indi-
cates species that might be important in
different geologic settings. Pay particular atten-
tion to dissolved species of possible sig-
nificance in assessing the reactivity of barium,
calcium, iron, magnesium, aluminum, and man-
ganese with injected wastes (see Table 6-5).
2. Select Laboratory Methods for Measuring
Parameters. Procedures for measuring basic
physical and chemical parameters are well-
established, although, as noted, different sources
may specify slightly different procedures.
Consult the U.S. EPA and/or the state
regulatory agency to identify preferred analytic
procedures for the specific parameters. If
specific procedures are not indicated, consult
the regulatory agency to determine whether es-
tablished procedures with which you are
familiar are acceptable. Tables 6-2 (reservoir
physical properties), 6-3 (methods for chemical
analyses), 6-4 (reservoir-fluid chemical proper-
ties), and 6-7 (subsurface microbial charac-
terization) list available analytic methods and
references with detailed descriptions of laboratory
procedures. Additional references may be
located using Table 6-9 (annotated bibliog-
raphy index).
3. Select Laboratory Methods for Testing Waste-
Reservoir Compatibility. Evaluate waste-
reservoir compatibility before undertaking any
other geochemical studies. In the worst case,
incompatibility may make deep-well injection
unfeasible (see the Wilmington, North Carolina,
case study, Section 7.5). In other instances, in-
compatibility can be handled by injecting in-
compatible waste streams separately or by
pretreating the waste stream to improve com-
patibility (see Texas Petrochemical case study,
Section 7.7). Section 6.3 discusses water-
reservoir interaction tests in some detail, and
Table 6-8 summarizes 14 such studies. The
Annotated Bibliography in Section 6.6 indicates
which references provide descriptions of
laboratory methods.
4. Select Laboratory Methods for Other Geochem-
ical Fate Studies. Laboratory waste-reservoir
interaction tests discussed in Section 6.3 may be
useful for other aspects of assessing the geochemi-
cal fate of injected wastes. Also, Section 6.4 discus-
ses a number of laboratory procedures for studying
chemical transformation processes.
6.2 Waste/Reservoir Characterization
Characterizing the wastes and the reservoir into which
they are injected requires measuring numerous
parameters. The following sections summarize key
parameters that may need to be measured in each of
the following areas:
• Waste stream (Section 6.2.1)
• Reservoir lithology (Section 6.2.2)
• Formation water (Section 6.2.3)
• Microbiology (Section 6.2.4)
6.2.1 The Waste Stream
A large number of parameters need to be measured
to characterize a waste stream. Table 6-1 lists more
than 30 parameters in five categories: (1) waste
volume, (2) physical properties, (3) chemical com-
position, (4) chemical reactivity, and (5) biological
characteristics.
• Waste volume is an important measure because of
limitations on the physical capacity of the reservoir
rock to accept wastes without unacceptable in-
creases in pressure.
• Physical properties affect the fbw of injected
wastes in the subsurface. For example, temperature,
density, and viscosity influence mixing processes in
the reservoir fluid. In the Belle Glade case study
(Section 7.4), the higher temperature and lower
relative density of the waste compared with those
of the reservoir fluid accelerated the upward
migration of the waste to a shallower aquifer after
the confining layer was breached. Also, the type
and form of solids and gases dictate the types of
pretreatment that can be used to reduce plugging
potential.
• Chemical composition influences reactivity (see
below) and the kinds of geochemical processes
that may occur. Whether the waste is classified
as hazardous or nonhazardous will determine
which sets of injection regulations apply.
• Chemical reactivity influences the design of the
injection well in several ways. Corrosivity and
120
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Table 6-1 Basic Parameters for Characterizing
Wastewater
I. Volume
Average daily flow rate
Duration and level of maximum flow rate
Maximum rate of change of flow rate
II. Physical Properties
Temperature range
Insoluble components: colloidal, settleable, floatable
Color and odor
Viscosity, density, compressibility
Radioactivity
Foamability
Dissolved oxygen
III. Chemical Composition
Known organic and inorganic components
Chemical oxygen demand, total carbon, extractables
pH, Eh, acidity, alkalinity
Oxidizing or reducing agents (sulfides)
Chloride ion
Hardness (calcium and magnesium)
Nitrogen and phosphorus
Surfactants
Chlorine demand
Total dissolved solids
Specific ions (see Tables 6-5 and 6-6)
Phenol
Grease and hydrocarbons
IV. Chemical Reactivity
Corrosiveness
Chemical stability
Reaction with injection-system components
Reaction with formation waters
Reaction with formation minerals
V. Biological Characteristics
Biochemical oxygen demand
Pathogenic bacteria
Chemical toxicity (aquatic life, bacteria, plants, humans)
Source: Adapted from Tables 5-1 and 5-5 in Warner and
Lehr(1977).
chemical stability (see Section 1.1.3) influence
the choice of materials. Incompatibility between
the waste and the injection system's
components, formation waters, and formation
minerals (which can cause precipitation, gas
bubbles, etc.) must be identified and corrected to
ensure proper functioning (see Section 3.3 and
Section 6.3). Watkins (1954) describes procedures
for corrosion testing.
• Biological characteristics will determine the
extent to which microbial action may cause
clogging and whether biodegradation may be
significant as injected wastes move through the
injection zone (see Sections 6.2.4 and 6.4.3).
Several EPA documents describe sampling methods
for wastes (deVera, 1980; Ford et al., 1984). A num-
ber of tests and methods are required to measure
specific waste-stream parameters. Comprehensive
compilations have been developed by Longbottom
and Lichtenberg (1982), Kopp and McKee (1983),
APHA (1985), and ASTM (1966; annual).
6.2.2 Reservoir Lithology
Both physical and chemical properties of the injec-
tion formation and confining layer must be measured
to characterize the lithology of the reservoir. Warner
and Lehr (1977) discuss field sampling methods for
rock from coreholes. Hewitt (1963) discusses
methods to evaluating water sensitivity of reservoir
rocks.
6.2.2.1 Physical Properties
Table 6-2 summarizes the physical properties of the
formation rock that must be identified before the fate of
wastes can be evaluated. Two types of parameters are
listed in this table—those which affect the chemical
reactivity of the host rock and those which affect the
physical flow of injected wastes. Warner and Lehr
(1977) discuss the significance of and methods for
measuring or estimating values for those physical
properties that primarily affect the physical flow of in-
jected wastes: porosity, bulk density, permeability,
compressibility, temperature, and stress. Chemical
reactivity of the host rock is influenced in part by the
physical parameters of texture (particle-size distribu-
tion) and specific surface area. Specific surface area
influences the amount of mineral surface available for
rock-fluid chemical interactions. Specific surface area
increases, from sand to silt to clay particle-size frac-
tions. Porosity and permeability influence chemical
reactivity by determining the extent to which fluids can
reach available surfaces. Geochemically, effective
porosity, the amount of interconnnected pore space
available for fluid transmission, is more important that
total porosity, because it is the effective porosity that
determines the amount of matrix surface available for
rock-fluid chemical reactions. Thus shale has a very
high surface area because it is composed of clay par-
ticles, but its low permeability means that little fluid can
flow into the rock to allow chemical reactions to
121
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Table 6-2 Physical Properties of Reservoirs
Important in Deep-Well Geochemical
Fate Assessment
Property
References on Analytic
Procedures
Rock
Texture (particle-size
distribution)
Porosity
Bulk density
Specific surface area
Permeability
Compressibility
Temperature
Stress
Klute, 1986
Warner and Lehr, 1977
Collins and Crocker, 1988
Klute, 1986
Klute, 1986
Warner and Lehr, 1977
Warner and Lehr, 1977
Warner and Lehr, 1977
Warner and Lehr, 1977
Reservoir Fluid
Viscosity
Warner and Lehr, 1977
Density (specific gravity) Warner and Lehr, 1977
Hem, 1970
Fluid pressure
Compressibility
Warner and Lehr, 1977
Kreitleretal., 1988
Warner and Lehr, 1977
take place. Section 3.1.4 discusses the importance
of particle-size distribution.
6.2.2.2 Chemical Properties
The mineralogy of the reservoir rock determines the
important chemical properties and strongly influences
the ion-exchange and adsorption capacity of the injec-
tion zone. Clay mineralogy is especially important be-
cause of the possible significance of clays in well
plugging (see Section 3.3.1) and the major role of clays
in adsorption. Section 3.1.4 (Reservoir Matrix) discus-
ses the geochemical significance of clays and other
minerals. Table 6-3 summarizes major laboratory
methods used to investigate reservoir rocks, including
mineral identification and measurement of adsorption
properties. Section 6.3 discusses laboratory methods
for determining waste-rock adsorption interactions.
6.2.3 Formation Water
Scalf et al. (1981), Berg (1982), and Barcelona et al.
(1985) describe methods for collecting ground-water
samples. In deep-well formations, samples must be
collected carefully to minimize the chemical changes
that can occur when the sample is brought to the
surface. For example, ferrous ion (Fe+2) in solution is
unstable in the presence of oxygen, and carbonates
and bicarbonates are particularly susceptible to equi-
librium shifts if samples are stored in plastic bottles
because of carbon-dioxide diffusion (Scalf et al.,
1981). Gases dissolved at deep-well pressures will
be lost when brought to the surface unless samples
are isolated from the atmosphere. Rose and Long
(1988) review ground-water sampling methods as
they apply to collecting dissolved oxygen. Sampling
for other gases would require similar measures.
6.2.3.1 Physical Properties
Warner and Lehr (1977) discuss the significance of
physical properties as they relate to reservoir fluids
in the injection zone and methods for measuring
and estimating these properties (see Table 6-2).
As noted in Section 6.2.1, these properties are
primarily significant in the mixing process between
formation water and injected waste. Fluid pressure,
combined with the physical parameters discussed
in Section 6.2.2 will determine pumping pressures.
Kreitler et al. (1988) describe methods for evaluat-
ing formation pressures (see also Section 3.1.5).
6.2.3.2 Chemical Properties
Table 6-4 lists chemical properties required to char-
acterize fully the range of formation fluids that may
be found in injection zones. Such properties as pH,
Eh, and total dissolved solids strongly influence the
geochemical processes that may occur in the deep-
well environment. Ostroff (1965) and Collins (1975)
discuss sampling of deep formation waters and
which constituents to analyze. Table 6-4 also lists
the section in this reference guide where a detailed
discussion of specific parameters can be found and
presents references on analytic procedures for each.
Note that not all the parameters listed in Table 6-4
must be determined for all formation fluids; for ex-
ample, oxygen and hydrogen radioisotope analysis
would be required only if the approximate age of the
water is desired (Kreitler et al., 1988).
The dissolved species that should be analyzed may also
vary depending on geologic conditions. The species
present and their concentrations will influence distribution-
of-species and precipitation-dissolution reactions with
122
-------
Table 6-3 Chemical Analysis Methods for Reservoir Rock
Property
Method
Sources on Methods
Mineralogy
Clay mineralogy
Ion exchange
Adsorption
X-ray diffraction
Chemical component analysis
Microscopic identification
(thin sections, heavy minerals)
Scanning electron microscopy
Cation-exchange capacity
Adsorption isotherms
Calorimetry
X-ray diffraction
Spectroscopy
UV-visible
Electron Spin Resonance (ESR)
Infrared (IR)
Kerr, 1959
Bentley et al., 1986
Carroll, 1970; Grim, 1968
Pageetal., 1986
Collins and Crocker, 1988
Royetal., 1987
Collins and Crocker, 1988
Carroll, 1970; Grim, 1968
Skoog, 1985
Mortland, 1970;Theng, 1974
Table 6-4 Chemical Properties of Reservoir Fluids Important in Deep-Well Geochemical Fate Assessment
Property
Sources for Analytic Procedures
Reference Guide Section No.
General Chemical Properties
PH
Eh
Conductivity
Alkalinity
Total dissolved solids
Barnes, 1964; Kreitleret al., 1988
Wood, 1976;ZoBell, 1946
Hem, 1970
Barnes, 1964
ASTM, annual; APHA, 1985;
Rainwater and Thatcher, 1960
3.1.1
3.1.2
3.1.3
3.1.1
3.1.3
Inorganic Parameters
Dissolved inorganic species3
Radioisotopes (oxygen, hydrogen)
ASTM, annual; APHA, 1985;
Rainwater and Thatcher, 1960
Kreitleret al., 1988
6.2.3.2
6.2.3.2
Organic Parameters
Dissolved organic carbon (DOC)
Total organic acids
Titrated organic alkalinity
Microbiota
Malcolm and Leenheer,1973
APHA, 1985; Kreitleret al., 1988
ibid.
American Petroleum Institute, 1965
3.1.4.3
6.2.3.2
6.2.4
aSee Table 6-5.
123
-------
injected wastes. Dissolved species in formation water
can be organized into four categories: (1) cations,
(2) anions, (3) gases, and (4) organic acids. Table 6-5
identifies the species that may be present. Within each
category, species are ranked according to abundance
(major, intermediate, and minor). The table also indi-
cates species of importance in evaluating the reactivity
of the reservoir fluids with injected wastes (i.e., when
precipitation reactions are a concern).
The presence of dissolved species in reservoir fluids is
site-specific. Available hydrogeochemical data for the
region and the formations of interest should be
evaluated before selecting the species to be analyzed
in samples from a potential injection zone. Table 6-6
shows chemical analyses, species analyzed, and con-
centrations measured in five published studies on
reservoir fluids in the injection zones; these studies
may provide some guidance in selecting species. The
data from Wilmington, North Carolina, are from coastal-
plain sediments where injection is no longer practiced
(see Section 7.5). The data from Pensacola and Belle
Glade, Florida, are from coastal-plain carbonate deposits
(see Sections 7.2 and 7.4). The Marshall sample was
taken from a Devonian limestone in the Illinois Basin
(see Figure 3-1 in Chapter Three). The Frio formation in
Texas is an unconsolidated sand that receives more in-
jected wastes than any other formation in the United
States.
Deep-well-injection-zone formation fluids are under
pressure, leading to difficulties measuring some of their
chemical parameters. For example, pH measurements
of deep basinal brines have always been considered
unreliable because CO2 degasses when the sample
depressurizes as it rises in the well bore or comes in
contact with the atmosphere (Kreitler et al. 1988). Kreit-
ler et al. (1988) discuss how titrated inorganic alkalinity
can be used to estimate pH in this situation. Sampling
techniques that capture any gases passing out of
Table 6-5 Classification of Dissolved Species in Deep-Well Formation Water
Abundance8
Intermediate
Minor
Cations
Anions
Gases
Organic Acids
Major
Sodium
Calciumb
Magnesium
Chloride
Bicarbonate6
Sulfate6
Carbon
dioxide
Acetate
Propionate
Silicab
Barium
Potassium
Strontium
Boron
Iron6
Aluminum
Manganese
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Copper
Lead
Lithium
Molybdenum
Nickel
Selenium
Zinc
Nitrate
Nitrite
Orthophosphate
Bromide
Iodide
Fluoride
Nitrogen
Hydrogen
Sulfide6
Methane
Butyrate
aAbundance classification criteria (mg/L): Major: 103-105; Intermediate: 101-103; Minor: <101.
bOf possible special significance in assessing reactivity with injected wastes.
124
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Table 6-6 Chemical Constituents of Formation Waters Analyzed in Studies Related to Deep-Well Injection
Constituent3
Depth (ft)
Temperature (°C)
Specific gravity
PH
Eh(mV)
Conductance
(%mho/cm3)
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
Butyrate
Total org. acids
Titrated org. alk.
Wilmington,
North Carolina
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
—
—
—
—
—
—
Pensacola,
Florida
1,430
35.2
—
7.4
-32
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
—
—
—
—
—
—
Belle Glade,
Florida
1,200
26.0
.008
8.1
—
—
—
136
—
—
36
19
140
140
—
—
—
540
1,680
1
—
—
—
<1
<1
<0.1
4
2
—
—
—
—
—
Marshall,
Illinois
2,395
—
—
9.1
-154
22,000
22,000
380
66
—
—
8
147
117
8,370
73
—
182
12,700
20
—
—
25
—
—
<0.1
—
—
—
—
—
—
—
Frio
Formation,
Texas
7,000
107
—
8.2
—
—
118,802
2,448
—
—
—
30
9,460
2,980
43,300
356
948
120
71,400
—
247
33
—
—
—
—
—
—
1,270
207
24
1,500
1,457
125
-------
Table 6-6 (Continued)
Constituent
Aluminum
Arsenic
Barium
Boron
Beryllium
Cadmium
Chromium
Cobalt
Copper
Iron (total)
Ferrous iron
Lead
Lithium
Manganese
Mercury
Molybdenum
Nickel
Selenium
Strontium
Zinc
Wilmington,
North Carolina
<1
<0.01
<1
—
—
<0.1
<0.1
<0.01
<0.1
2
—
<0.01
<1
<-,
0.01
<0.01
<0.01
<0.01
19
<0.1
Pensacola, Belle Glade, Marshall,
Florida Florida Illinois
— — 0.09
<0.1 — <0.05
— — 1
5 — —
— — <0.01
— — <0.03
— — 0.26
— — <0.03
<0.1 — <0.03
— — 0.12
2 — —
<0.1 — <0.03
<1 - -
<0.1 — 0.14
— — —
— — 0.07
— — <0.05
— — —
22 — —
<0.1 — <0.03
Frio
Formation,
Texas
89
474
—
—
—
_
—
999
—
—
—
_
—
—
—
—
405
—
aUnless otherwise noted, all values are in mg/L.
Sources: Wilmington, NC., Leenheer et al. (1976); Pensacola, FL, Goolsby (1972); Belle Glade, FL, Kaufman et al.
(1973); Marshall, IL, Roy et al. (1989); Frio Formation, TX., Kreitler et al. (1988).
solution during depressurization should be used, and
the gaseous phase of the sample should be analyzed
quantitatively. The types of gases present in brines
are also important indicators of microbiological ac-
tivity, discussed in detail below.
6.2.4 Microbiology
Dunlap et al. (1977), Bitton and Gerba (1984), and
Ghiorse and Wilson (1988) describe subsurface sam-
pling methods for microbiological characterization, which
requires information on (1) types of microorganisms
(species, morphologic groups, etc.), (2) biomass, and
(3) viability (how much of the biomass is alive or
engaged in metabolic activities). Table 6-7 summarizes
the major techniques for characterizing the subsurface
and indicates which properties they may be able to iden-
tify. Note that most techniques must be used in combina-
tion with others. For example, cultures must be examined
microscopically as well, and epifluorescence light micros-
copy is most frequently combined with chemical,
radioisotope, and dye reduction techniques. Rosswall
(1973) and Costerton and Colwell (1979) contain detailed
discussions of various study techniques, and Webster et
al. (1985) provide a more recent discussion of several
methods specifically applicable to subsurface samples.
Section 6.4.2 discusses some laboratory methods for
evaluating the biodegradability of hazardous wastes and
simulating biodegradatbn in the deep-well environment.
6.3 Waste-Reservoir Interaction Tests
After measuring or estimating the values for the relevant
waste, rock, and fluid parameters, the researcher can
perform waste-reservoir interaction tests to: (1) identify
possible incompatibilities between reservoir components
and wastes to be injected, (2) identify chemical interac-
tions, and (3) provide data for predicting the fate of in-
jected wastes. Specific procedures for performing
126
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Table 6-7 Methods for Subsurface Microbial Characterization
Properties Measured
Technique
Species/Morph.
Mass
Viability
Microscopic Analyses
Scanning electron microscope (SEM)a
Transmission electron microscope (TEM)
Epifluorescence light microscopy
Cultures
Aufwuchs methods
Nutrient cultures
Chemical Analyses
Muramicacid
Lipid phosphates
Adenosine triphosphate (ATP)
Guanosine triphosphate (GTP)/ATP ratio
Inorganic metabolic substrates/products
Radioisotope Analyses
3H-thymidine in DMA
3H-glucose
Dye-reduction Analysis
Dye reduction
aOf limited value because of generally low population densities.
Source: Adapted from Ghiorse and Balkwill (1985).
interaction tests described in the literature vary con-
siderably but can be grouped into two major types:
batch tests and flowthrough tests.
Batch tests are performed by mixing wastes and
reservoir materials. The materials are mixed in a series
of reactors, which may be subjected to temperatures
and pressures that simulate the deep-well environ-
ment. At regular intervals, the reactors are opened in
sequence and the fluids analyzed. When waste and
reservoir fluids are mixed, the presence and type of
precipitates may be the main concern; when injection
fluid is mixed with reservoir rock, adsorption or dissolu-
tion reactions may be of primary interest, and changes
in the concentration of species being adsorbed or dis-
solved species may be measured. (Sections 5.2.2 and
6.4.1 describe the process for developing adsorption
isotherms). When neutralization reactions are of in-
terest, measurements continue until the mixture
shows no further change in pH. Either disaggregated
(crushed) or undisturbed cores may be used in batch
tests. When undisturbed cores are used, special pro-
cedures should be used to ensure complete satura-
tion of the core material by the fluid being tested.
Probably the single best source of information on
batch-test procedures for estimating adsorption is
Roy etal. (1987).
Flowthrough tests also are used to study interac-
tions between fluids and solids. The solid may be an
undisturbed core or packed columns intended to
simulate subsurface conditions. In either case the
same core is used throughout, and the injected fluid
is monitored at the outflow end at specified time
127
-------
intervals to observe changes in chemistry. In adsorp-
tion experiments, equilibrium is obtained when the
outflow concentration equals the inflow concentra-
tion. If precipitation-dissolution reactions occur, pres-
sure changes caused by clogging or increased
permeability may be monitored in addition to chemi-
cal changes.
Table 6-8 summarizes information on 14 waste-
reservoir interaction tests reported in the literature. It
lists the type of test, type of waste, geologic-formation
lithology, and, where indicated, the duration and the
temperature and pressure conditions. Most of the refer-
ences in this table describe equipment and procedures
for performing the tests. Roy et al. (1989) (batch-test)
and Collins and Crocker (1988) (flowthrough tests) are
useful, recent sources.
The following issues should be considered when
selecting a laboratory method for evaluating interac-
tions between wastes and reservoir materials:
• The results of any method will contain uncertainties
created by the sample chosen (which may not be
representative of the injection zone), and possible
alteration of in situ properties caused by shaping.
Furthermore, because such experiments are
usually performed for hours or days, only those
reactions which reach equilibrium quickly will be
measured. Reactions taking years to reach
equilibrium will not be measured.
• Tests must simulate temperatures and pressures in
the injection zone, unless preliminary tests show
that these parameters do not significantly affect the
processes of interest. For example, Elkan and
Horvath (1977) performed preliminary tests of
microbiological activity at pressures similar to those
in the injection zone being simulated and found no
significant difference between activity at the
elevated pressure and that at atmospheric
pressure. Subsequent experiments were con-
ducted at atmospheric pressure.
• Results from tests using simulated sand cores or
simulated waste solutions have lower confidence
levels than those using actual cores and waste
streams.
• Batch experiments using disaggregated material
are very likely to overestimate adsorption rates
because of the larger surface area created by
disaggregation. Batch experiments using undis-
turbed cores are very likely to yield better results,
but they still will not simulate subsurface conditions
as effectively as flowthrough experiments in
undisturbed cores.
• Flowthrough experiments on subsurface cores at
simulated temperatures and pressures will probably
yield the best results, although the uncertainties
listed above still apply (i.e., whet her.the sample is
representative of the injection zone and whether
the experiment's duration allows full equilibrium to
be reached).
6.4 Geochemical Processes
This section describes laboratory procedures related to
the study of (1) adsorption, (2) hydrolysis, and (3) bio-
degradation. The study of other processes such as acid-
base equilibria, dissolution-precipitation, neutralizaton,
and complexation all involve batch or flowthrough tests
as described in the previous section.
6.4.1 Adsorption Isotherms
Adsorption isotherms (see Section 5.2.2.1) can be
measured using either batch (Donaldson and Johan-
sen, 1973; Donaldson et al., 1975) or flowthrough
experiments (Collins and Crocker, 1988). Either pro-
cedure requires a series of measurements determin-
ing over time (at constant temperature) the changes
in concentration of a solution with a known starting
concentration. Equilibrium adsorption is reached when
there is no significant change in concentration of the
substance between measurements. Roy et al. (1987)
recommend using a rate of change in solute con-
centration of less than 5 percent per 24-hour time in-
terval. The amount adsorbed at equilibrium (usually
expressed as micrograms/gram [fig/g] solid) can be
calculated by dividing the amount adsorbed (begin-
ning concentration minus the final concentration) by
the weight of the adsorbing solid. Plotting the equi-
librium adsorption value at different concentrations
and fitting the data into the appropriate equation form
(see Section 5.2.2.1) allow adsorption to be calcu-
lated at other concentrations. Measuring adsorption
isotherms at two or more temperatures allows the
heat-of-adsorption to be estimated, which may be
valuable when interpreting thermodynamic mechan-
isms. Collins and Crocker (1988) describe proce-
dures to estimate heat of adsorption from such data.
The previous section (6.3.1) discusses the factors
that should be considered in selecting procedures for
measuring adsorption.
6.4.2 Hydrolysis
The formulas for calculating rates of hydrolysis (KH) are
discussed in Section 5.2.4. If rate constants for a par-
ticular substance cannot be found in the literature, they
128
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Table 6-8 Summary of Waste-Reservoir Compatibility/Interaction Studies
Waste Type
Time Temp Pressure
Formation (days) (°C) (MPa)
Source
Batch— Fluids
Acidic, organic
(diluted)
49 Organic
compounds
Subsurface bacteria 3 20 0.1-27.6
culture
Various bacterial 2-8 37-56 0.1
cultures
Elkan and
Horvath,
1977
Grula and
Grula, 1976
Batch — Disaggregated
Acidic, inorganic
Alkaline, organic
Acidic, organic
Acidic, ferric
chloride
Cresol, sodium
borate
St. Peter sandstone 15 25-55 0.1-11.7
Potosi dolomite
Proviso si Itstone 155- 52a 10.8a
Brine (Devonian) 230a
Floridan limestone — — 5.07
Dolomite 0.25 43 6.89
Bentonite — 250 0.1
Roy et al.,
1989
Goolsby,
1972
Hower e,t
al, 1972
Apps et
al., 1988
Batch— Undisturbed
Various organics
Various organics
Cottage Grove — 60 20.3
sandstone
Cottage Grove — 38-93 20.7
sandstone
Donaldson
and Johan-
sen, 1973
Donaldson
Flo wthrough— Column
Unspecified
Acidic, organic
Miocene sand — — —
Cretaceous sand 80 20 0.1
(simulated)
Hower et
al., 1972
Elkan and
Horvath,
1977
129
-------
Table 6-8 (Continued)
Waste Type
Formation
Time Temp
(days) (°C)
Pressure
(MPa)
Source
Organic acids,
formaldehyde
Acidic pickling
liquor
Phenol (in
simulated brine)
Flowthrough—Undisturbed
Acidic (steel)
Mt. Simon sandstone — ? 0.1 Bayazeed
and
Donaldson,
1973
Poorly consolidated
sands
Dolomitic sandstone,
dolomite, quartzite
Frio sand
6-18
20
40
3.4
13.8
38-60 24.1
Leenheer et
al., 1976
Ragone et
al., 1978
Collins and
Crocker, 1988
Unpublished data provided by W. R. Roy, Illinois State Geological Survey.
must be determined by laboratory tests. As noted in
Section 2.3.3, the effects of high ionic strength on
hydrolysis rates are difficult to predict, so tests should
simulate deep-well salinity conditions. Hydrolysis rate
constants should probably be measured in simulated
deep-well conditions to determine how significantly other
environmental factors affect hydrolysis rates. Mill et al.
(1982) describe procedures for measuring hydrolysis
rate constants. Suffet et al. (1981) contain some addi-
tional suggestions for procedures not included in Mill et
al. (1982).
6.4.3 Biodegradation
As with adsorption, biodegradation may be tested in
the laboratory using either batch (Grula and Grula,
1976) or flowthrough (Elkan and Horvath, 1977) experi-
ments. Pritchard and Bourquin (1984) review the use of
microcosms when studying interactions between pol-
lutants and microorganisms. Table 6-8 summarizes
temperatures and pressures that have been used to
simulate biodegradation under deep-well conditions.
The batch experiments performed by Grula and Grula
(1976) used enrichment cultures containing single or-
ganic compounds. After suitable incubation (typically
24 hours), four or five serial transfers to media of the
same composition were made to enhance the selection
of microbes with degradative capacity for the specific
compound.
Elkan and Horvath (1977) have conducted the most
sophisticated laboratory simulations of microbial
degradation of injected wastes. Flowthrough experi-
ments using sand-packed columns inoculated with
bacterial populations from the injection zone were
used to simulate conditions at the Wilmington, North
Carolina, case study (Section 7.5). Fluids that simu-
lated the composition and concentration of the actual
waste were injected through the columns for 40 to 80
days under a range of experimental conditions. Al-
though the study provided many useful insights into
microbial degradation processes, it was not able to
simulate their full range of activity. Specifically, all
the model experiments had at least one factor that
inhibited methanogenesis, which was known to occur
in the injection zone.
Where a well is already operating, backflow tests can
be used to evaluate biodegradation (see Section 7.1
and case studies in Sections 7.2 and 7.3). The types
of biodegradation occurring can be inferred by test-
ing samples for inorganic degradation byproducts
(i.e., methane, hydrogen sulfide, and nitrogen).
6.5 Quality-Assurance/Control
Procedures
EPA regulations require a quality-assurance and
quality-control (QA/QC) plan that covers all aspects of
a no-migration demonstration. Minimal requirements
for a QA/QC program are described in U.S. EPA
(1976) and detailed procedures for laboratory analyses
130
-------
are presented by Booth et al. (1979); other
regulatory agencies may require additional or dif-
ferent procedures. In addition to using standard and
approved methods, the researcher must also define
the principles and objectives of the QA/QC plan and
identify the personnel responsible for implementing
it.
Deep-well geochemical-fate modeling will present spe-
cial challenges for QA/QC planning. The researcher
must document original derivations of thermodynamic
and kinetic data and obtain all experimental data using
analytical methods acceptable to EPA and/or the ap-
propriate regulatory agency. All procedures should be
calibrated and referenced to those of the National
Bureau of Standards (NBS), the U.S. EPA, or other ac-
ceptable body. Deviations from acceptable procedure
must be fully documented, with demonstrations that al-
ternatives yield the same- or better-quality results than
those currently accepted by the U.S. EPA.
6.6 Annotated Bibliography
6.6.1 How To Use this Bibliography
Table 6-9 lists the references in this bibliography by
topic. The annotations briefly describe each document
and may identify chapters or sections most likely to be
of interest. The references cited may be obtained from
(1) the publisher, (2) libraries, (3) the originating trade
association, (4) the originating government agency, (5)
the Department of Commerce's National Technical In-
formation Service (NTIS), or (6) the author.
6.6.2 Annotations
American Petroleum Institute (API). 1965. Recom-
mended Practices for Biological Analysis of Subsur-
face Injection Waters. RP 38. API, 1220 L St. NW,
Washington, D.C. 20005.
Describes methods for (1) identifying microor-
ganisms in a water sample by microscopic examina-
tion, general bacterial counts, and sulfate-reducing
bacteria counts, and (2) determining effectiveness of
chemicals for treating injection water to prevent the
growth of sulfate-reducing bacteria.
American Public Health Association (APHA). 1985.
Standard Methods for the Examination of Water and
Wastewater, 16th edition. 1,268 pp. AHPA, 1015 Fif-
teenth Street NW, Washington, D.C. 20005.
Most pertinent sections include: Part 100, General
Introduction (laboratory apparatus, precision,
accuracy, quality control; Part 200, Physical Ex-
amination (color, conductivity, solids); Part 300,
Determination of Metals; Part 400, Determination
of Inorganic Nonmetallic Constituents; Part 500,
Determination of Organic Constituents (organic
carbon, BOD, COD); and Part 900, Microbiologi-
cal Examination of Water.
American Society for Testing and Materials (ASTM).
Annual Books of ASTM Standards. Water and Environ-
mental Technology, Volumes 11.01 and 11.02 (Water).
ASTM, 1916 Race St., Philadelphia, Pennsylvania
19103.
American Society for Testing and Materials (ASTM).
1966. Manual on Industrial Water and Industrial Waste-
water, 2nd edition. ASTM, 1916 Race St., Philadelphia,
Pennsylvania 19103.
Part I covers general use of industrial water and
problems of sampling and analysis and Part II sets
forth ASTM standards. Of particular interest is
D1256-61 (Scheme for Analysis of Industrial Wastes
and Industrial Waste Water). Analytical methods for
a range of inorganic and organic chemical
parameters are presented.
Apps, J. 1988. Current Geochemical Models to Predict
the Fate of Hazardous Wastes in the Injection Zones of
Deep Disposal Wells. Draft Report LBL-26007.
Lawrence Berkeley Laboratory, Berkeley, CA 94720.
Comprehensive report on the state of the art in
geochemical modeling of interactions between haz-
ardous wastes and the injection reservoir. Contains
eight major sections: (1) an introduction to the EPA
regulations covering no-migration petitions for deep-
well injection of hazardous wastes and how the peti-
tions relate to geochemical modeling; (2) discussion
of the reactions that must be modeled given the
chemical conditions expected in the injection zone;
(3) the equations of state that must be used; (4) the
availability of thermodynamic data; (5) the availability
of geochemical modeling computer codes; (7) criteria
affecting the satisfactory chemical modeling of
waste injection; and (8) conclusions and recommen-
dations. U.S. EPA (1989) contains a summary of
this report.
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
Engineering, March 9-10, 1988. LBL-24337. Lawrence
Berkeley Laboratory, Berkeley, CA 94720, pp. 79-82.
Focuses on chemical aspects of deep-well injec-
tion of hazardous wastes, including: (1) an over-
view of types of models for predicting fate and
deficiencies in available models; (2) a comparison
131
-------
Table 6-9 Topical Index to Annotated Bibliography
Topic
Sources
Waste Stream
Hazardous waste characterization
API, 1965; APHA, 1985; ASTM, annual; ASTM, 1966; Bayzeed and Donaldson,
1973; Berg, 1982; deVera, 1980; Ford et al., 1984; Kopp and McKee. 1983;
Longbottom and Lichtenberg, 1982; Malcolm and Leenheer, 1973- Mill et al
1982; Warner and Lehr, 1977; Watkins, 1954
Resevolr Rock
Physical characteristics
Chemical characteristics
Mineralogy
ASTM, annual; Bentley et al.. 1986; Collins and Crocker, 1988; Klute, 1986;
Kreitler et al.. 1988; USGS, var. dates; Warner and Lehr, 1977
ASTM, annual; Page et al., 1986; Mortland, 1970; USGS, var. dates; Warner and
Lehr, 1977; ZoBell, 1946
Bentley et al.. 1986; Carroll, 1970; Grim, 1968; Hewitt, 1963; Kerr, 1959; Klute,
1986; Mortland, 1970; Theng, 1984
Resevoir Fluids
Physical characteristics
Chemical characteristics
ASTM, annual; Collins, 1975; Ostroff, 1965; USGS, var. dates; Warner and Lehr
1977
APHA, 1985; ASTM, annual; Apps et al., 1988; Barnes, 1964; Berg, 1982; Collins,
1975; Dunlap et al.. 1977; Hem, 1970; Kopp and McKee, 1983; Kreitler, 1988;
Malcolm and Leenheer, 1973; Ostroff, 1965; Rainwater and Thatcher, 1960; Scalf
etal., 1981; USGS, var. dates; Warner and Lehr, 1977; Watkins, 1954; Wood
1976; ZoBell, 1946
Other
Waste-reservoir interactions
Biological characterization
Biodegradation
Hydrolysis
Sampling
Quality assurance/control
Apps, 1988; Collins and Crocker, 1988; Donaldson and Johnasen, 1973; Elkan
and Horvath, 1977;Goolsby, 1972; Hewitt, 1963; Howeretal., 1972; Kaufman et
al., 1982; Ostroff, 1965; Pritchard and Bourquin, 1984; Ragone et al 1978' Roy
et al., 1987; U.S. EPA, 1989; Warner and Lehr, 1977
API, 1965; APHA, 1985; Barnes, 1972; Bayzeed and Donaldson, 1973; Bitton and
Gerba, 1984; Costerton and Colwell, 1979; Dunlap et al., 1977; Elkan and
Horvath, 1977; Ghiorse and Balkwill, 1985; Ghiorse and Wilson, 1988; Pritchard
and Bourquin, 1984; Rosswall, 1973; Webster, 1985
Bitton and Gerba, 1984; Elkan and Horvath, 1977; Ghiorse and Wilson, 1982;
GrulaandGrula, 1976; Lyman etal., 1982, Pritchard and Bourquin, 1989
Lyman et al., 1982; Mill et al., 1982; Suffet et al., 1981
Barcelona et al., 1985; Berg, 1982; Bitton and Gerba, 1984; DeVera, 1980;
Dunlap et al., 1977; Ford et al., 1984; Ghiorse and Wilson, 1988; Rainwater and
Thatcher, 1960; Scalf et al., 1981; USGS, var. dates; Warner and Lehr, 1977
APHA, 1985; Berg, 1982; Booth et al., 1979; U.S. EPA, 1976
132
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of the laboratory simulation of the evolution of Gulf
Coast brines using EQ3/6 with actual brines; and
(3) the results of laboratory experiments studying
the interactions between bentonite clay and a
simulated waste of sodium borate and cresol. U.S.
EPA (1989) contains a summary of this report.
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.
Covers all aspects of ground-water sampling
(parameter selection, well placement and construc-
tion, monitoring-well design, well development, and
recommended sampling protocols.
Barnes, I. 1964. Field Measurement of Alkalinity and
pH. U.S. Geological Survey Water Supply Paper 1525-
H. 17pp.
Describes detailed procedures for accurate meas-
urement of alkalinity and pH in the field. pH meter
readings in the field may be in error as much as
0.5 pH units when the pH of a sample differs
greatly from the reference (buffer solutions).
Barnes, I. 1972. Water-Mineral Reactions Related to
Potential Fluid-Injection Problems. In Symposium on
Underground Waste Management and Environmental
Implications, Houston, Texas, T. D. Cook, ed. Am.
Assn. Petr. Geol. Mem. 18, pp. 294-297.
Discusses thermodynamic aspects of precipitation-
dissolution reactions in natural ground waters. No
specific data are reported on waste-reservoir inter-
actions.
Bayazeed, A. F., and E. C. Donaldson. 1973. Subsur-
face Disposal of Pickle Liquor. U.S. Bureau of Mines
Report of Investigations 7804.30 pp.
Surveys underground injection of steel-processing
waste pickle liquor. Includes (1) data on analyses of
waste from several steel companies; (2) a detailed
case history of injection into the Mt. Simon sandstone
near Gary, Indiana; (3) the results of laboratory ex-
periments on water sensitivity (permeability changes
due to changes in salinity) of Mt. Simon and Cottage
Grove sandstones; and (4) results of laboratory flow-
through experiments using simulated hydrochloric
acid pickling liquor.
Bentley, M. E., R. T. Kent, and G. R. Myers. 1986. Site
Suitability for Waste Injection, Vickery, Ohio. In Proc. of
the Int. Symp. on Subsurface Injection of Liquid Wastes,
New Orleans. National Water Well Association, Dublin,
Ohio 43017. pp. 330-354.
Site evaluation of a waste injection site in the Mt.
Simon Sandstone. Describes results of laboratory
tests to characterize core materials (x-ray diffraction,
microscopic examination of thin sections of cores to
identify minerals and characterize porosity, per-
meability and porosity), and field tests (drill-stem and
pressure-fall-off analysis).
Berg, E. L. 1982. Handbook for Sampling and Sample
Preservation of Water and Wastewater. EPA 600/4-82-
029. NTIS PB83-124503.
Particularly relevant chapters are: (1) General Con-
siderations for a Sampling Program; (4) Statistical
Approach to Sampling; (6) Sampling Industrial
Wastewaters; (9) Sampling of Ground and Drinking
Water; (12) Sampling, Preservation and Storage
Considerations for Trace Organic Materials; (15)
Sample Control Procedures and Chain of Custody;
(16) Quality Assurance; and (17) Sample Preserva-
tion. Note that this report supercedes a report by the
same title by Moser et al. (1976), EPA/600/4-80-
029.
Booth, R. L., et al. 1979. Handbook of Analytical
Quality Control in Water and Wastewater Laboratories.
EPA/600/4-79-019, NTIS PB 297 451.
Details quality control procedures for laboratories
that analyze water and wastewater samples.
Bitton, G., and C. P. Gerba (eds.). 1984. Groundwater
Pollution Microbiology. Wiley-lnterscience, New York.
Paper by McNabb and Mallard presents methods for
obtaining uncontaminated subsurface samples for
microbiological analysis.
Carroll, D. 1970. Clay Minerals: A Guide to their X-Ray
Identification. GSA Special Paper No. 126. Geological
Society of America, Box 9140, Boulder, Colorado
80301.
Not obtained for review.
Collins,'A. G. 1975. Geochemistry of Oilfield Waters.
Elsevier, New York, 496 pp.
Particularly relevant chapters include: (2) Sampling of
Oilfield Waters (dissolved gases, unstable con-
stituents, pH/Eh flow sampling chamber), (3) Analysis
of Oilfield Water for Physical Properties and Inorganic
133
-------
Chemical Constituents, (4) Interpretation of
Chemical Analyses of Oilfield Waters, (5) Sig-
nificance of Some Inorganic Constituents and
Physical Properties of Oil Field Waters, (6) Or-
ganic Constituents in Oilfield Waters, (12) Com-
patibility of Oil Field Waters.
Collins, A. G., and M. E. Crocker. 1988. Laboratory
Protocol for Determining Fate of Waste Disposed in
Deep Wells. EPA-600/8-88-008, NTIS PB88-166061.
Describes laboratory procedures for: (1) core
analysis, (2) brine analysis, (3) dynamic flow-
through system that simulates the interactions of
hazardous organic wastes with injection zone
rock, and (4) static waste/rock interaction tests
that simulate longer-term degradation processes.
Protocol testing resulted in some data on the ad-
sorption of phenol and 1,2,-dichloroethane in
simulated subsurface conditions for the Frio
sandstone, and data from earlier adsorption ex-
periments using the Cottage Grove sandstone are
presented (see Table 5-4). U.S. EPA (1989) con-
tains a summary of this report.
Costerton, J. W. and R. R. Colwell (eds.). 1979. Na-
tive Aquatic Bacteria: Enumeration, Activity, and
Ecology. ASTM/Special Tech. Pub. 695. (See ASTM
[1965] for address).
Papers presented at a symposium sponsored by
ASTM in June, 1977. Contains five papers of
methods for direct enumeration of aquatic bac-
teria, five papers on chemical indices of aquatic
bacterial populations, and six papers on metabolic
potentials of aquatic bacterial populations as indi-
cated by activity measurements.
deVera, E. R. 1980. Samplers and Sampling Proce-
dures for Hazardous Waste Streams. EPA 600/2-80-
018. NTIS PB80-135353.
Describes procedures for collecting, handling, stor-
ing, and recording samples of hazardous wastes.
Various sampling devices are discussed, with em-
phasis on developing a composite liquid-waste
sampler (the coliwasa).
Donaldson, E. C., and R. T. Johansen. 1973. History of
a Two-Well Industrial Waste Disposal System. In Sym-
posium on Underground Waste Management and Ar-
tificial Recharge, J. Braunstein, ed. Pub. No. 110, Int.
Assn. of Hydrological Sciences, pp. 603-621.
Case history of a facility injecting separate acidic
and alkaline organic waste streams in Texas (see
case study in Section 7.7 and Table 5-4).
Donaldson, E. C., M. E. Crocker, and F. S. Manning.
1975. Adsorption of Organic Compounds on Cottage
Grove Sandstone. BERC/RI-75/4, Bartlesville Energy
Research Center, Bartlesville, Oklahoma.
Presents data on the results of adsorption of nine
organic compounds on the Cottage Grove
sandstone at 3,000 psi and two temperatures
(100° and 150°F). See Table 5-4 for summary of
data results.
Dunlap, W. J., J. F. McNabb, M. R. Scalf, and R. L.
Cosby. 1977. Sampling for Organic Chemicals and
Microorganisms in the Subsurface. EPA 600/2-77-
176. NTIS PB272679.
Describes methods for obtaining samples of
ground-water and earth solids that are not con-
taminated by near-surface microbes. Procedures
described are limited to a depth of about 25 feet
below the surface in compact alluvial formations.
Elkan, G., and E. Horvath. 1977. The Role of
Microorganisms in the Decomposition of Deep-Well-
Injected Liquid Industrial Wastes. NSF/RA-770102
NTIS PB 268 646.
Presents results of (1) studies to characterize pre-
injection and post-injection microbial populations in
injection zones at the Monsanto plant, Florida (see
Section 7.2) and Wilmington, North Carolina (see
Section 7.5), and (2) laboratory apparatus and pro-
cedures for simulating microbial degradation of in-
jected wastes.
Ford, P. J, P. J. Turina, and D. E. Seely. 1984. Char-
acterization of Hazardous Waste Sites—A Methods
Manual: Vol II: Available Sampling Methods, 2nd edi-
tion. EPA 600/4-84-07. Vol III. Available Laboratory
Analytical Methods. Available from EPA Cincinnati.
Section 3.4 of Vol. II describes procedures for
purging and sampling monitoring wells. Vol. Ill out-
lines detailed methodology suitable for hazardous
waste analysis and is organized by media and
compound.
Ghiorse, W. C., and D. L. Balkwill. 1985. Microbiologi-
cal Characterization of Subsurface Environments. In
Ground Water Quality, C. H. Ward, W. Giger, and P. L.
McCarty, eds. Wiley Interscience, New York, pp. 386-
401.
Reviews available methods for characterizing sub-
surface microorganisms.
134
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Ghiorse, W. C., and J. T. Wilson. 1988. Microbial
Ecology of the Terrestrial Subsurface. Adv. Appl.
Microbiol. 33:107-172.
Section III (Characterization of Microorganisms
and Their Activities in Subsurface Environments)
includes subsections on sampling and methods
for detection, enumeration, and metabolic activity.
Goolsby, D. A. 1972. Geochemical Effects and Move-
ment of Injected Industrial Waste in a Limestone Aquifer.
In Symposium on Underground Waste Management
and Environmental Implications, Houston, Texas, T. D.
Cook, ed. Am. Assn. Petr. Geol. Mem. 18, pp. 355-368.
Reports studies of geochemical interactions be-
tween an acidic organic waste and a carbonate
injection zone (see Monsanto case study, Sec-
tion 7.2).
Grim, R. E. 1968. Clay Mineralogy, 2nd edition. Mc-
Graw-Hill, New York.
Chapters of particular interest include: (5) X-Ray
Diffraction Data, (7) Ion Exchange, (8) Clay Water
System, (9) Clay Mineral-Organic Reactions, and
(11) Optical Properties.
Grula, M. M., and E. A. Grula. 1976. Feasibility of
Microbial Decomposition of Organic Wastes under
Conditions Existing in Deep Wells. BERC/RI-76/6.
Bartlesville Energy Research Center, Bartlesville,
Oklahoma.
Presents results of aerobic biodegradation experi-
ments on 50 compounds in ten organic groups
(mono- and di-carboxylic acids, aldehydes and
ketones, amino acids, alcohols, mono- and di-
amines, aliphatic nitro compounds, nitriles,
aromatic compounds, and miscellaneous) under
simulated deep-well temperatures (50° to 70°C)
and pressures (100 atm).
Hem, J. D. 1970. Study and Interpretation of the
Chemical Characteristics of Natural Water, 2nd edi-
tion. U.S. Geological Survey Water Supply Paper
1473.
Provides data and discussion for more than 60
constituents and their properties that are included
in water analyses for which sufficient data exist to
consider sources from which each is generally
derived, most probable form of elements and ions
in solution, solubility controls, expected concentra-
tion ranges, and other chemical factors. Also dis-
cusses statistical techniques for analyzing water
quality data.
Hewitt, C. H. 1963. Analytical Techniques for Recog-
nizing Water Sensitive Rocks. J. of Petroleum Tech-
nology 15:813-818.
Describes the following techniques for identifying
water-sensitive rocks: (1) flowthrough permeability
tests, (2) x-ray diffraction, (3) physical swelling
tests, and (4) microscopic examination of thin sec-
tions. Presents typical analyses of reservoir rocks
that are not sensitive, that are water-sensitive due
to swelling clays, and that are water-sensitive due
to particle plugging.
Hower, W. F., R. M. Lasater, and R. G. Mihram.
1972. Compatibility of Injection Fluids with Reservoir
Components. In Symposium on Underground Waste
Management and Environmental Implications, Hous-
ton, Texas, T. D. Cook, ed. Am. Assn. Petr. Geol.
Mem. 18, pp. 287-293.
Discusses clay-mineral sensitivity, presents
results of laboratory flow tests in water-sensitive
sand, and presents ferric-chloride-dolomite
interaction tests.
Kaufman, M. I., D. A. Goolsby, and G. L. Faulkner.
1973. Injection of Acidic Industrial Waste into a
Saline Carbonate Aquifer: Geochemical Aspects. In
Symposium on Underground Waste Management
and Artificial Recharge, J. Braunstein, ed. Pub. No.
110, Int. Assn. of Hydrological Sciences, pp. 526-
551.
Reports results of studies of geochemical interac-
tions between acidic organic wastes and a car-
bonate injection formation in Belle Glade, Florida
(see Case Study, Section 7.4).
Kerr, P. F. 1959. Optical Mineralogy, 3rd edition. Mc-
Graw Hill, New York. 442 pp.
Provides comprehensive coverage of optical
methods for the microscopic identification of
minerals.
Klute, A. (ed.). 1986. Methods of Soil Analysis, Part
1—Physical and Mineralogical Methods, 2nd edition.
ASA Monograph 9. American Society of Agronomy,
677 S. Segoe Rd., Madison, Wisconsin 53711, 1,188
pp.
Contains 50 chapters covering a range of physical
and mineralogical methods. Many of the methods
for physical characterization can be used to char-
acterize geologic materials.
135
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Kopp, J. F., and G. D. McKee. 1983. Methods for
Chemical Analysis of Water and Wastes. EPA 600/4-
79-020, revised March 1983, NTIS PB84-128677.
This third edition contains the chemical analytical
procedures used in U.S. EPA laboratories for ex-
amining ground and surface waters, domestic and
industrial waste effluents, and treatment process
samples. Provides test procedures for measuring
physical inorganic and selected organic con-
stituents and parameters. Supercedes report by
the same title dated 1979 (NTIS PB 297 686).
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.
Section 8.3 describes methods for analyzing
anions (chloride, bromide, sulfate, ammonia, and
iodide), and isotopically analyzing oxygen and
hydrogen, and organic acids and organic alkalinity.
Section 8 also contains water quality analyses of
about 850 samples from the Frio formation in
Texas, the most widely used injection zone for
hazardous wastes.
Leenheer, J. A., R. L Malcolm, and W. R. White. 1976.
Physical, Chemical and Biological Aspects of Subsur-
face Organic Waste Injection near Wilmington, North
Carolina. U.S. Geological Survey Professional Paper
987.
Comprehensively reports on results of field and
laboratory waste-aquifer reactivity studies at the
Wilmington, North Carolina, injection facility (see
case study, Section 7.5).
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.
Describes test procedures for 15 groups of or-
ganic chemicals, and includes an appendix defin-
ing procedures for determining the detection limit
of an analytic method. The test procedures in this
manual are cited in Tables 1C (organic chemical
parameters) and 1D (pesticide parameters) in 40
CFR 136.3(a).
Lyman, W. J., W. F. Reehl, and D. H. Rosenblatt,
eds. 1982. Handbook of Chemical Property Estima-
tion Methods: Environmental Behavior of Organic
Compounds. McGraw-Hill, New York.
The most relevant chapters for deep-well injection
are Chapter 6 (Rate of Hydrolysis) by Harris, and
Chapter 9 (Rate of Biodegradation) by Scow.
Malcolm, R. L., and J. A. Leenheer. 1973. The Useful-
ness of Organic Carbon Parameters in Water Quality
Investigations. In Proc. of the Inst. of Env. Sciences
1973 Annual Meeting, Anaheim, CA, April 1-6, pp. 336-
340. Available from J. A. Leenheer, USGS MS 408,
Box 25046, Federal Center, Denver, Colorado, 80225.
Describes methods for sampling and analysis for
dissolved organic carbon (DOC) and suspended
organic compounds (SOC). Also discusses the
relationship of DOC and SOC to other organic in-
dices (BOD, COD, and TOC).
Mill, T., W. R. Mabey, D. C. Bomberger, T. -W. Chou,
D. G. Hendry and J. H. Smith. 1982. Laboratory
Protocols for Evaluating the Fate of Organic Chemicals
in Air and Water. EPA-600/3-82-022, NTIS PB83-
150888.
Particularly relevant protocols for deep-well injec-
tion are described in Chapter 4 (Hydrolysis in
Water) and Chapter 8 (Sorption of Organic on
Sediments). Each chapter contains procedures for
preliminary screening and detailed tests.
Mortland, M. M. 1970. Clay-Organic Complexes and
Interactions. Adv. Agron. 22:75-117.
Discusses methods for studying clay-organic
bonding mechanisms, different types of bonding
mechanisms, and the nature of some clay-organic
complexes and reactions.
Ostroff, A. G. 1965. Introduction to Oilfield Water
Technology. Prentice Hall, Englewood Cliffs, New
Jersey, 412pp.
Particularly relevant chapters for industrial-waste
injection include: (2) Analysis of Water (sampling
methods, determination of components), (3) Scales
and Sludges Deposited from Water, (4) Water and
Corrosion, (6) Water Treatment Microbiology, and
(13) Water for Injection (compatibility tests, water-
sensitive formations).
Page, A. L., R. H. Miller, and D. R. Keeney, eds.
1986. Methods of Soil Analysis, Part 2—Chemical
136
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and Microbiological Properties, 2nd edition. ASA
Monograph 9, American Society of Agronomy, 677
S. Rd., Madison, Wisconsin 53711.1,159 pp.
Contains 54 chapters covering methods for
analyzing chemical and microbiological properties
of soils.
Pritchard, P. H., and A. W. Bourquin. 1984. The Use
of Microcosms for Evaluation of Interactions Be-
tween Pollutants and Microorganisms. Adv. Microbial
Eco/. 7:133-215.
Comprehensively reviews the use of microcosms
(establishment of a physical model or simulation
of part of an ecosystem) for studying biodegrada-
tion. Topics covered include: role of microcosms in
environmental risk assessment, typical microcosm
design features, qualitative and quantitative ap-
proaches to using microcosms, and field calibration.
Ragone, S. E., R. S. Riley, and R. J. Dingman. 1978.
Hydrochemistry and Hydrodynamics of Injecting an
Iron-Rich Pickling Liquor into a Dotomitic Sandstone—
A Laboratory Study. J. Res. U.S. Geol. Survey 6(1):1-9.
Describes development of a high-pressure
permeability-testing apparatus for dynamic flow-
through waste/reservoir interaction experiments and
presents results of experiments in which acidic
waste pickling liquor containing high concentrations
of iron salts was injected into cores of quartzite,
sandstone, and dolomite.
Rainwater, F. H. and L. L. Thatcher. 1960. Methods
for Collection and Analysis of Water Samples. U.S.
Geological Survey Water Supply Paper 1454, 301
PP-
Pertinent sections are: (A) Collection of Samples
(site selection, frequency, equipment and sam-
pling instructions); (B) Handling of Water Samples
before Analysis; (C) Analysis of Water Samples
(types of methods, choice of analytical method);
and (D) Analytical Procedures (specific proce-
dures for over 40 inorganic water parameters).
Rose, S., and A. Long. 1988. Monitoring Dissolved
Oxygen in Ground Water: Some Basic Considera-
tions. Ground Water Monitoring Review8C\):93-97.
Paper reviewing the geochemical significance of
dissolved oxygen in ground water and sampling
methods for dissolved gases.
Rosswall, T., ed. 1973. Modem Methods in the Study
of Microbial Ecology. Bulletins from the Ecological
Research Committee, Swedish Natural Science Re-
search Council, Stockholm, 17.
Includes about 80 papers and short communica-
tions presented at a symposium held in Up-
psala, Sweden, in 1972. Pertinent sessions
include: (2) Techniques for the Observation of
Microcosms in Soil and Water; (3) Isolation and
Characterization of Microorganisms; (4) Techni-
ques for the Determination of Microbial Activity in
Relation to Ecological Investigations; (5) Estima-
tion of Microbial Growth Rates Under Natural
Conditions; (6) Model Systems; (7) Mathematical
Models and Systems Analysis in Microbial Ecol-
ogy. Also includes summary of a panel discussion
on problems of assessing the effect of pollutants
on microorganisms.
Roy, W. R., S. C. Mravik, I.G . Krapac, D. R. Dicker-
son, and R. A. Griffin. 1989. Geochemical Interactions
of Hazardous Wastes with Geological Formations in
Deep-Well Systems. Environmental Geology
Notes 130. Illinois State Geological Survey,
Champaign, Illinois. [An earlier version of this report
by the same title was published in 1988 by the Haz-
ardous Waste Research and Information Center,
Savoy, Illinois].
Includes: (1) a description of laboratory proce-
dures for batch-type waste-rock-brine interaction
tests at simulated subsurface temperature and
pressure conditions; (2) data on geochemical in-
teractions at different temperatures and pressures
between two types of hazardous waste (acidic
and alkaline) with material from two injection-
zone formations and one confining formation that
occur in the Midwest (Mt. Simon sandstone,
Potosi dolomite, and Proviso siltstone; and (3) a
comparison of the empirical data with predictions
using two aqueous geochemical codes (WATEQ2
and SOLMNEQF). U.S. EPA (1989) contains a
detailed summary of this report.
Roy, W. R., I. G. Krapac, S. F. J. Chou, and R. A. Grif-
fin. 1987. Batch-Type Adsorption Procedures for Es-
timating Soil Attenuation of Chemicals. Draft Technical
Resource Document (TRD), EPA/530-SW-87-006-F.
NTIS PB87-146155. [The final TRD, titled Batch-Type
Procedures for Estimating Soil Adsorption of Chemi-
cals, is scheduled for publication in 1990]
Provides a comprehensive report on batch-test
procedures for estimating soil adsorption. Includes
chapters on selection of soil:solution ratios for
ionic and nonionic solutes, determination of equi-
libration time, construction of adsorption isotherms,
137
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selection of adsorption equations, application of
batch-adsorption data and laboratory procedures
for generating adsorption data. Literature on the
effects of temperature, pH, ionic strength, phase
separation, method of mixing, and soihsolution
ratios on adsorption were also reviewed.
Scalf, M. R., J. F. McNabb, W. J. Dunlap, and R. L.
Cosby. 1981. Manual of Ground-Water Quality Sam-
pling Procedures. EPA 600/2-81-160. NTIS PB82-
103045.
Specific methods are covered in the following
chapters: (5) Construction of Monitoring Wells,
(6) Collection of Ground Water Samples, (7) Sam-
pling Subsurface Solids. Appendices of interest in-
clude: (A) Summary of Procedures Based on
Parameters of Interest and (B) Sampling of Low
Density Immiscible Organics).
Skoog, D. A. 1985. Principles of Instrumental
Analysis, 3rd Edition. Saunders College Publishing,
Philadelphia, Pennsylvania.
Presents a comprehensive text on laboratory
methods for measurement of physical and chemi-
cal properties of materials.
Suffet, I. H., C. W. Carter, and G. T. Coyle. 1981.
Test Protocols for the Environmental Fate and Move-
ments of Toxicants. In Proceedings of a Symposium
of the AOAC, G. Zweig, and M. Beroza, eds. As-
sociation of Official Analytical Chemists, Washington,
D.C.
Suggests refinements to laboratory protocols for
measuring hydrolysis rates outlined in Mill et al
(1982).
Theng, B. K. G. 1974. The Chemistry of Clay-Organic
Reactions. Adam Hilger Ltd., London.
Reports on the chemistry of clay-organic reac-
tions, with special emphasis on the use of infrared
spectroscopy.
U.S. Environmental Protection Agency, 1989. As-
sessing the Geochemical Fate of Deep-Well-Injected
Hazardous Waste: Summaries of Recent Research.
U.S. EPA 625/6-89-025b.
Presents in a standardized summary format the
following research papers: (1) Apps (1988),
(2) Apps et al. (1988), (3) Collins and Crocker
(1988), (4) Roy et al. (1989), and (5) Strycker and
Collins (1987).
U.S. Environmental Protection Agency. 1976. Mini-
mal Requirements for a Water Quality Assurance
Program. EPA/440/9-75-010, NTIS PB 258 807.
Presents a guide for planning and developing a
quality assurance program. Part II includes a typi-
cal Memorandum of Understanding (MOD) on
quality assurance procedures between an EPA
regional office and a state. Other parts cover over-
all requirements, basic elements to be imple-
mented immediately, and basic elements to be
implemented in the future.
U.S. Geological Survey. Various Dates. Techniques
of Water-Resources Investigations of the United
States Geological Survey.
Presents an ongoing compilation of water
resource investigation methods. Some chapters of
interest include: Book 5, Chapter A2 (Determina-
tion of Minor Elements in Water by Emission
Spectroscopy), Chapter A3 (Methods for Analysis
of Organic Substances in Water), Chapter D2
(Guidelines for Collection and Field Analysis of
Groundwater Samples for Selected Unstable Con-
stituents).
Warner, D. L., and J. H. Lehr. 1977. An Introduction
to the Technology of Subsurface Wastewater Injec-
tion. EPA 600/2-77-240. NTIS PB 279 207.
Arguably the best single reference document
covering all aspects of the design, construction, and
operation of waste-injection wells. Chapter 6
(Wastewater Characteristics) discusses sampling in
general and the parameters that should be con-
sidered in characterizing wastes.
Watkins, J. W. 1954. Analytical Methods of Testing
Waters to be Injected into Subsurface Oil-Productive
Strata. U.S. Bureau of Mines Report of Investigations
5031.29 pp.
The main section of interest covers corrosion tests.
It also describes procedures for measuring ten
parameters of potential significance in assessing
corrosivity of a liquid (dissolved oxygen, free carbon
dioxide, hydrogen sulfide, pH, total and dissolved
iron, alkalinity and carbonate stability, hardness,
chlorides, residual chlorides, and turbidity).
138
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Webster, J. J., G.J. Hampton, J.T. Wilson, W.C.
Ghiorse, and F.R. Leach. 1985. Determination of
Microbial Cell Numbers in Subsurface Samples.
Ground Water 23:17-25.
Describes procedures for several methods (ATP
measurements and AOINT counts) for determin-
ing microbial cell numbers in subsurface samples
and reports results of applying the methods to
samples from Oklahoma and Texas.
Wood, W. W. 1976. Guidelines for Collection of Field
Analysis of Ground-Water Samples for Selected Un-
stable Constituents. Chapter D2 of Techniques of
Water-Resource Investigations of the United States
Geological Survey. Available from U.S. Geological
Survey, Books and Open-File Reports Section,
Federal Center, Box 25425, Denver, Colorado
80225.
Covers field methods for sampling and field
analysis of specific conductance, temperature,
pH, carbonate and bicarbonate, Eh, and dissolved
oxygen.
ZoBell, C. E. 1946. Studies on Redox Potential of
Marine Sediments. Am. Ass. Petr. Geol. Bull.
30:477-513.
Describes procedures for colorimetric and
electrometric measurement of Eh (redox potential)
of sedimentary materials.
139
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CHAPTER SEVEN
CASE STUDIES OF DEEP-WELL INJECTION OF INDUSTRIAL WASTE
This chapter discusses how field studies can be used
in geochemical fate assessment (Section 7.1) and in-
cludes six cases of deep-well-injection facilities,
documenting the geochemistry of the hazardous and
other industrial wastes injected as reported in the
literature. Each case study is organized in the same
format, with section headings as follows:
• Injection Facility Overview describes the type of
facility, its current status, and the characteristics of
the injected wastes and presents a brief history of
injection and monitoring activities, including distance
traveled by the waste.
• Injectton/Confining-Zone Lithotogy and Chemistry
provides information on the geology and chemistry of
the injection zone formation fluids.
• Chemical Processes Observed briefly describes
the types of interactions and major physical effects
that have been observed at the site and evaluates
their significance.
Table 7-1 summarizes information about each study,
including chapter reference, location of the well, lithol-
ogy of the injection zone, waste characteristics, major
geochemical processes observed, and sources of in-
formation.
Other case-study compilations of industrial-waste injec-
tion have been prepared by Donaldson (1964),
Donaldson et al. (1974), and Reeder et al. (1977). The
first two summarize information on source and nature of
waste, geology, surface equipment, and well completion
and operation for 15 companies. Reeder et al. (1977)
describe about a dozen studies in a similar format, or-
ganized by EPA region. Companies and precise loca-
tions are not specified in these studies, and none
particularly emphasize geochemical interactions be-
tween injected waste and the reservoir formation. Never-
theless, these compilations provide useful information on
geologic characteristics of injection zones, waste
pretreatment, and injection-well operating characteristics.
7.1 Use of Field Studies in Geochemical
Fate Assessment
Field studies are an important complement to
geochemical modeling, as discussed in Chapter
Five, and to laboratory studies, as discussed in
Chapter Six. Two ways to investigate interactions
between injected wastes and reservoir material are
(1) direct observation of the injection zone and over-
lying aquifers using monitoring wells and (2) back-
flushing the injected waste. In both instances,
samples of the fluids in the zone are collected at in-
tervals to characterize the nature of geochemical
reactions and to track changes over time.
7.1.1 Monitoring Wells
Monitoring wells drilled into the injection zone at
selected distances and directions from the injection
well allow direct observation of formation water char-
acteristics and the interactions that occur when the
waste front reaches the monitoring well. When
placed near the injection well in the aquifer above
the confining layer, monitoring wells can detect the
upward migration of wastes caused by casing or
confining-layer failure. Foster and Goolsby (1972)
describe detailed methods for constructing monitor-
ing wells.
Monitoring wells have several advantages: time-
series sampling of the formation over extended
periods is easy and the passage of the waste front
can be observed precisely. Disadvantages are cost
and the potential for upward migration of wastes if
monitoring well casings fail. A monitoring well at the
Monsanto plant had to be plugged when unneutral-
ized waste reached it because of fears that the
casing would corrode (see Section 7.2.1). The three
Florida case studies (Sections 7.2, 7.3, and 7.4) and
the North Carolina case study (Section 7.5) illustrate
the usefulness of monitoring wells.
141
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Table 7-1 Summary of Case Studies
Location Lithology Wastes
Processes
Observed
Section Sources
Florida
Pensacola Limestone Nitric acid Neutralization
(Monsanto) Inorganic salts Bacterial
Organic compounds denitrification
7.2 Goolsby, 1971,1972
Faulkner and Pascale, 1975
Pascale and Martin, 1978
Elkan and Horvath, 1977
Dean, 1965
Barraclough, 1966
Goolsby, 1971
Willis etal., 1975
Pensacola Limestone
(American
Cyanamid)
Belle Glade Carbonate
Aery lonit rile
Sodium salts
(nitrate,
sulfate
thiocyanate)
Hot acid
Organic plant
wastes
Bacterial 7.3
denitrification
No retardation
of thiocyanate
ions
Neutralization 7.4
Bacterial
sulfate reduction
Methane
production
Ehrlich et al., 1979
Vecchioli et al., 1984
Kaufman et. al., 1973
Kaufman and McKenzie, 1975
McKenzie, 1976
Garcia-Bengochea and Vernon, 1970
North Carolina
Wilmington Sand
Silty sand
Limestone
Organic acids
Formaldehyde
Methanol
Neutralization 7.5
Dissolution-
precipitation
Complexation
Adsorption
Bacterial sulfate
and iron
reduction
Methane
production
DiTommaso and Elkan,1973
Leenheerand Malcolm, 1973
Peek and Heath, 1973
Leenheer et al., 1 976a,b
Elkan and Horvath, 1977
Willis etal., 1975
Illinois
Tuscola
Dolomite
Hydrochlorite
acid
Neutralization
Dissolution
CO2 gas
production
7.6 Kamath and Salazar, 1986
Panagiotopoulos and Reid, 1986
Brower et al., 1989
Texas
Not
specified
Miocene
sand
(1) Organic acids
Organic
compounds
(2) Alkaline salts
Organic
compounds
Precipitation
Adsorption
(inferred)
7.7
Donaldson and Johansen, 1973
142
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7.1.2 Back flushing of Injected Wastes
Backf lush ing of injected wastes can also be a good
way to observe waste/reservoir geochemical interac-
tions. Injected wastes are allowed to backflow (if for-
mation pressure is above the elevation of the
wellhead) or are pumped to the surface. Backflowed
wastes are sampled periodically (and reinjected
when the test is completed); the last sample taken
will have had the longest residence time in the injec-
tion zone. Keely (1982) and Keely and Wolf (1983)
describe this technique for characterizing contamination
of near-surface aquifers and suggest using logarithmic
time intervals for chemical sampling. The three Florida
studies (Sections 7.2, 7.3 and 7.4) all present results
from backflushing experiments.
The advantages of backflushing are reduced cost com-
pared with that of monitoring wells and reduced sam-
pling time (sampling takes place only during the test
period). Disadvantages include less precise time- and
distance-of-movement determinations and the need to
interrupt injection and to have a large enough area for
backflushed fluid storage before reinjection.
7.2 Case Study No. 1: Pensacola,
Florida (Monsanto)
7.2.11njection-Facility Overview
Monsanto operates one of the world's largest nylon
plants on the Escambia River about 13 miles north of
Pensacola, Florida. The construction, operations,
and effects of the injection-well system at this site
have been extensively documented by the U.S.
Geological Survey in cooperation with the Florida
Bureau of Geology. Pressure and geochemical ef-
fects are reported by Goolsby (1972), Faulkner and
Pascale (1975), and Pascale and Martin (1978). Ad-
ditional microbiological data are reported by Willis et
al. (1975) and Elkan and Horvath (1977). Major
chemical processes observed at the site include
neutralization, dissolution, biological denitrification,
and methanogenesis. The geochemical fate of or-
ganic contaminants in the injected wastes, however,
has not been reported.
The waste is an aqueous solution of organic
monobasic and dibasic acids, nitric acid, sodium and
ammonium salts, adiponitrile, hexamethylenediamine,
alcohols, ketones, and esters (Goolsby, 1972). The
waste also contain cobalt, chromium, and copper, each
in the range of 1 to 5 mg/L. Waste streams with dif-
ferent characteristics, produced at various locations
in the nylon plant, are collected in a large holding
tank; this composite waste is acidic. The specific
characteristics of the waste varied somewhat as a
result of process changes, (e.g., after 1968 more
organic acids and nitric acid were added). Until
mid-1968, wastes were partially neutralized by
pretreatment. After that, unneutralized wastes were
injected. No reason was reported for suspending
treatment. Goolsby (1972) reports pH measurements
ranging from a high of 5.6 in 1967 (at which time the
pH was raised before injection by adding aqueous
ammonia) to a low of 2.4 in 1971, and Eh ranging
from +300 mV in 1967 to +700 mV in 1971. The
chemical oxygen demand in 1971 was 20,000 mg/L
(see Section 3.1.2 for a discussion of this
parameter).
Monsanto began injecting wastes into the lower lime-
stone of the Floridan aquifer in 1963. In mid-1964, a
second well was drilled into the formation about
1,000 ft southwest of the first. A shallow monitoring
well was placed in the aquifer above the confining
layer about 100 ft from the first injection well, and a
deep monitoring well was placed in the injection
zone about 1,300 ft south of both injection wells. The
deep monitoring well (henceforth referred to as the near-
deep monitoring well) was plugged with cement in 1969
(see below). In late 1969 and early 1970, two additional
deep monitoring wells were placed in the injection for-
mation, 1.5 miles south-southeast (downgradient) and
1.9 miles north-northwest (upgradient) of the site. From
1963 to 1977, about 13.3 billion gallons of waste were in-
jected. During the same period, injection pressures
ranged from 125 to 235 psi. Figure 7-1 shows the loca-
tion of all wells as of 1977; since then, a third injection
well has been added (Haberfeld, 1989a).
Ten months after injection of neutralized wastes began,
chemical analyses (see Section 7.2.3) indicated that
dilute wastes had migrated 1,300 ft to the nearest deep
monitoring well. Injection of unneutralized wastes
began in April 1968. Approximately 8 months later, un-
neutralized wastes reached the near-deep monitoring
well, indicating that the neutralization capacity of the
injection zone between the injection wells and the
monitoring well had been exceeded. At this point, the
monitoring well was plugged with cement from bottom
to top because operators were concerned that the
acidic wastes could corrode the steel casing and
migrate upward (Goolsby, 1972). The rapid movement
of the waste through the limestone indicated that most
of it migrated through a more permeable section, which
was about 65 ft thick. By mid-1973, 10 years after
injection began, a very dilute waste front arrived at the
south monitoring well, 1.5 miles away. As of early
1977, there was no evidence that wastes had reached
the upgradient monitoring well. The shallow monitoring
well remained unaffected during the same period.
143
-------
Figure 7-1 Location of Three Monitoring and Two Injection Wells, Monsanto Facility
(Pascale and Martin, 1978).
87° 19'
30° 38' -
30° 33' -
144
-------
Increases in permeability caused by limestone dis-
solution approximately doubled the injection index
(the amount of waste that can be injected at a
specified pressure). As of 1974, the effects of the
pressure created by the injection were calculated to
extend more than 40 miles radially from the injection
site (Faulkner and Pascale, 1975). An updip move-
ment of the freshwater/saltwater interface in the
injection-zone aquifer, which lies less than 20 miles
from the injection wells, was also observed (see
Figure 7-2).
7.2.2 InJection/Confining-Zone Lithology and
Chemistry
The tower limestone of the Floridan aquifer is used as
the injection zone (at 1,400 to 1,700 ft), and the
Bucatunna clay member of the Byram formation (about
220 ft thick) serves as the confining layer. Figure 7-3
shows the stratigraphy of the area, and Figure 7-4
shows the focal stratigraphy and the monitoring well in-
stallations. The formation water in the injection zone is
a highly saline (11,900 to 13,700 mg/L total dissolved
solids [TDS]) sodium-chloride solution. The Eh of
samples collected from two monitoring wells located
in the injection formation ranged from +23 to -32 mV,
indicating reducing conditions in the injection zone
that would favor anaerobic biodegradation. Table 6-6
in Chapter Six contains additional data on the
chemistry of the Floridan aquifer formation water.
The injection zone contains about 7,900 mg/L chloride,
but less than 20 miles northeast of the injection site,
chloride concentrations are less than 250 mg/L Under
natural conditions, water in the injection zone moves
slowly south-southwestward toward the Gulf of Mexico,
where it is assumed to discharge about 100 miles off-
shore. The preinjection hydraulic gradient was about
1.3 ft/mile (see Figure 7-2).
7.2.3 Chemical Processes Observed
As a result of dissolution of the limestone by the partly
neutralized acid wastes, calcium concentrations more
than doubled in the near-deep monitoring well 10
months after injection started in 1963 (Goolsby, 1972).
In early 1966, however, they dropped to background
levels (about 200 mg/L), possibly in response to
biochemical decomposition of the waste. In September
Figure 7-2 Hydrogeology of the Lower Limestone of the Floridan Aquifer in Northwest Florida
(Goolsby, 1972).
8 6° JO'
31*00'
S\ANTA\ ROSA OKALOOSA
£jfJJJ-. Areo where soltne water from lower
polentiometric surface of Ihe lower
limestone above meon sea level .pre-
1963. Contour interval, 20 feet.
Isochlor shows inferred chloride
concentration (milligrams per liter)
within upper part of the lower 11 mist on*
Flondon aquifer moves upward ond
mixes with fresh water in upper Floridort
aquifer under natural condilions
FouU . doshed where inferred
145
-------
Figure 7-3 Generalized North-South Geologic Section Through Southern Alabama and
Northwestern Florida (Goolsby, 1972).
500'
SEA
LEVEL
500'
IOOO'
I500'
2000'
2500'
3000'
3500
(Alter Borroclough, 1966)
ABOUT 100 MILES TO DISCHARGE AREA *•
50-15'
,' CCMJNTY ;
X. I SANTA ROS*
-\ s «-
* ) iiNjEciioN snr
ir
o
—I 500'
SEA
LEVEL
500'
IOOO'
1500'
2000'
25OO
3000'
LOCATION MAP
*-ABOUT 75 MILES TO RECHARGE AREA-
3500'
1968, after about 300 million gallons of the acidic,
unneutralized waste had been injected, the calcium
concentration began to increase again. An abrupt in-
crease in calcium to 2,700 mg/L accompanied by a
decrease in pH to 4.75 in January 1969 led to the
decision to plug the near-deep monitoring well.
In an attempt to find out how fast the waste was react-
ing with limestone, a 3-hour backflushing experiment,
in which waste was allowed to flow back out of the
injection well, yielded some unexpected results. The in-
crease in pH of the neutralized waste could not be fully
accounted for by the solution of limestone as deter-
mined from the calcium content of the backflushed liq-
uid; the additional neutralization apparently resulted
from reactions between nitric acid and alcohols and
ketones in the original waste induced by increased
pressure in the injection zone compared to surface
conditions (Goolsby 1971).
The lack of nitrates (which were present at levels of 545
to 1,140 mg/L in the waste) in the near-deep monitoring
well, combined with the presence of nitrogen gas, indi-
cated that degradation by denitrifying bacteria had taken
place (Goolsby, 1972). Backflushing shortly before inject-
ing unneutralized wastes confirmed denitrification.
Nitrate concentrations decreased rapidly as the
backflushed waste was replaced by formation water.
Similar backflushing experiments conducted after un-
neutralized wastes were injected, however, provided no
evidence of denitrification, indicating that microbial ac-
tivity was suppressed in the portion of the zone contain-
ing unneutralized wastes.
Elkan and Horvath (1977) performed a microbiological
analysis of samples taken from the north and south
deep monitoring wells in December 1974, about 6
months after the dilute waste front had reached the
south well. Both denitrifying and methanogenic bacteria
were observed. The lower numbers and species diver-
sity of organisms observed in the south monitoring well
compared with those in the north well indicated sup-
pression of microbial activity by the dilute wastes.
Chemical analyses of the north and south monitoring
wells were published for the period March 1970 to
March 1977 (Pascale and Martin, 1978). Between
September 1973 and March 1977 bicarbonate con-
centrations increased from 282 mg/L to 636 mg/L and
dissolved organic carbon increased from 9 mg/L to
146
-------
Figure 7-4 Monsanto Injection Facility Hydrogeologic Cross-Section (Faulkner and
Pascale, 1975).
DEEP INJECTION
MONITOR WELL WELL B
(PLUGGED)
NORTH DEEP
MONITOR WELI
SOUTH DEEP
MONITOR WELI
SAND-AND-GRAVEL
AQUIFER
IFRESH-WATER SUPPLY)
NSACOLA CLAYV
JPPER LIMESTONE
"LORIDAN AQUIFER
:HLORIDE. 350-450 mq/x
BUCATUNMA CLAY MEMBER
THE BYRAM FORMATION
(CONFINING BED
LOWER LIMESTONE
FLORIDAN AQUIFER
CHLORIDE 5,800-10,000 mg/L
(INJECTION ZONE )
SHALE AND CLAY
\\N\\\\
1800' -
EXPLANATION
CEMENT GROUT
CEMENT PLUG
NOTE
I- BOTTOM 20 FEET OF INNER CASING IS
STAINLESS STEEL IN THE TWO
INJECTION WELLS AND THE
TWO ACTIVE DEEP MONITOR WELLS.
2-ALL CASING GROUTED BOTTOM TO
TOP.
3-LINE OF SECTION SHOWN IN
FIGURE I.
HORIZONTAL NOT TO SCALE
47 mg/L. These increases were accompanied by an in-
crease in the dissolved-gas concentration and a dis-
tinctive odor like that of the injected wastes. The pH,
however, remained unchanged. During the same
period, dissolved methane increased from 24 mg/L to 70
mg/L, indicating increased activity by methanogenic bac-
teria. The observation of denitrification in the near-deep
monitoring well and methanogenesis in the more distant
south monitoring well fit the redox-zone biodegradation
model discussed in Section 5.2.3.3 (see also Table 5-5).
Significant observations made at this site are: (1) organic
contaminants (as measured by dissolved organic
carbon) continue to move through the aquifer even when
acidity has been neutralized, and (2) even neutralized
wastes can suppress microbial populations.
7.3 Case Study No. 2: Pensacola,
Florida (American Cyanamid)
7.3.11njection-Facility Overview
American Cyanamid Company operates a plant near
Milton, Florida, which lies about 12 miles northeast of
Pensacola and about 8 miles east of the Monsanto
plant discussed in Section 7.2. Chemical changes
caused by the injection of acidic wastes from this plant
have been reported by Ehrlich et at. (1979) and
Vecchioli et al. (1984), with the former citation provid-
ing the most complete information on the site. This
case study illustrates the complexity of assessing the
geochemical fate of mixed wastes. Acrylonitrile was
detoxified by biological reduction, whereas sodium
thiocyanate remained unaltered.
147
-------
The facility combines acidic waste streams from
various plant operations in a holding pond where they
are mixed and aerated. The waste is pumped from the
pond and neutralized with sodium hydroxide. The
neutralized wastes are treated with alum to flocculate
suspended solids and then passed through mixed-media
filters. A small amount of hydrogen peroxide solution
(amount unspecified) is added before filtration to inhibit
microbial growth on the filters. The pretreated waste that
is injected contains high concentrations of sodium
nitrate, sodium sulfate, sodium thiocyanate (an inorganic
cyanide compound), and various organic compounds, in-
cluding acrytonitrile (a listed hazardous waste—see
Table 4-10). The average pH of the waste is 5.8;
average chemical oxygen demand is 1,690 mg/L
A primary injection well and a standby well are situated
about 1,500 ft apart. A shallow monitoring well is located
near the primary injection well in the upper limestone
Floridan aquifer that overlies the confining Bucatunna
clay. Two deep monitoring wells in the injection zone are
located 1,000 ft southwest, and 8,170 ft northeast of the
primary injection well. Figure 7-5 shows the locations of
the injection and monitoring wells.
Waste injection began in June 1975, and waste was
first detected in the downgradient southwest deep
monitoring well about 260 days later. To analyze the
waste's physical and chemical properties after injec-
tion, the primary injection well was allowed to backflow
into a holding pond for 5 days in November, 1977. This
waste was sampled periodically (and reinjected when
the test was completed). About 4 years after injection
began (mid-1979), dilute waste arrived at the standby
injection well 1,560 ft south of the primary well.
7.3.2 Injection/Confining-Zone Lithologyand
Chemistry
The injection well is in the same area as the Monsanto
well, so the geology and native-water chemistry are
very similar to that described in Section 7.2. Figure 7-6
shows the stratigraphy of the immediate area and dis-
tances between the injection and monitoring wells. The
tower limestone of the Floridan aquifer is used as the
injection zone (1,230 to 1,440 ft), and the confining
Bucatunna clay is about 165 ft thick. TDS levels range
from 12,000 to 12,700 mg/L, with chloride ion con-
centrations of 6,700 mg/L. The pH ranges from 7.3 to
7.6, and temperature, from 30° to 32°C. Table 6-6 con-
tains additional data on the chemistry of the formation
water. Caliper and ftowmeter tests made in the injec-
tion wells suggest that the waste moves almost ex-
clusively within the top 18 m (55 ft) of the tower
limestone. As discussed in Section 7.2.2, the preinjec-
tton ground-water flow direction is south-southwest
(see Figure 7-2).
7.3.3 Chemical Processes Observed
The Eh of the injected waste dropped rapidly from +40
mV to -80 mV in the first 40 hours after injection began
and remained at about -80 mV thereafter. Denitrifying
bacteria detoxified the acrytonitrile by mineralizing the
compound, breaking it down into bicarbonate and am-
monia. The nitrates were degraded to nitrogen gas.
The backflow test described in Section 7.3.1 produced
data indicating that these transformations were about
90% complete within 82 ft of the injection well and vir-
tually 100% complete within 328 ft. These results are
an example of a biodegradation-disperston curve (see
Figure 2-2 in Chapter Two). Denitrrfying-bacteria den-
sities increased from traces (101 organisms/100 mL in
the native ground water) to large populations (107 to
10 organisms/100 mL) in injected wastes that had
been in the aquifer for several days.
Sodium thiocyanate (NaSCN) was first detected in
the closest monitoring well (1,000 ft away) 260 days
after injection began. Ammonium ions (a reaction
product of biomineralization) did not appear as a
contaminant until 580 days after initial injection. This
delay was probably the result of ion exchange or
other adsorption processes and may be an example
of an adsorption-dispersion curve (see Figure 2-2 in
Chapter Two). Because sodium thiocyanate in the
waste remained unchanged during its movement
through the injection zone, it was used to detect the
degree of mixing that took place between the waste liq-
uid and native water in an observation well. Thus the
appearance of sodium thiocyanate as well as an in-
crease in chemical oxygen demand in the standby well
4 years after injection began signaled the arrival of
wastes at that location.
This case study is interesting in that one hazardous
waste (acrytonitrile) was quickly rendered nonhazardous
after injection, whereas another (sodium thiocyanate)
showed no evidence of decomposition during the dura-
tion of the study. The implication for geochemical fate as-
sessment is that research should focus on the
compounds likely to be most resistant to decomposition
and/or immobilization, since they will be the ones most
critical in demonstrating containment in a no-migration
petition.
7.4 Case Study No. 3: Belle Glade,
South-Central Florida
7.4.11njection-Facility Overview
The Belle Glade site, located southeast of Lake
Okeechobee in south-central Florida (see Figure 7-7),
illustrates some of the problems that can develop with
acidic-waste injection when carbonate rock is the
148
-------
Figure 7-5 Location of the American Cyanimid Injection Site and Monitoring Wells
(Ehrlich et al., 1979).
confining layer. Contributing factors to the contamination
of the aquifer above the confining zone were the dissolu-
tion of the carbonate rock and the difference in density
between the injected wastes and the formation fluids.
The injected waste was less dense than the ground
water because of its tower salinity and higher tempera-
ture (Kaufman et al., 1973).
The injected fluids include the effluent from a sugar mill
and the waste from the production of furfural, an al-
dehyde processed from the residues of processed
sugar cane. The waste is hot (about 75° to 93°C);
acidic (pH 2.6 to 4.5); and has high concentrations of
organics, nitrogen, and phosphorus (Kaufman and
McKenzie, 1975). The waste is not classified as haz-
ardous under 40 CFR 261, and the well is currently
regulated by the State of Florida as a nonhazardous
injection well (Haberfeld, 1989b). The organic carbon
concentration exceeds 5,000 mg/L.
The well was originally cased to a depth of 1,495 ft,
and the zone was left as an open hole to a depth of
1,939 ft. The depth of the zone has been increased
twice (see later discussion). Seasonal injection (fall,
winter, and spring) began in late 1966; the system
was inactive during late summer. Injection rates
ranged from 400 to 800 gallons per minute, and
wellhead injection pressures ranged from 30 to 60
psi. By 1973 injection had become more or less con-
tinuous. From 1966 to 1973, more than 1.1 billion
gallons of waste were injected (Kaufman et al.,
1973).
At the time injection began, a shallow monitoring well
was placed 75 ft south of the injection well in the upper
149
-------
Figure 7-6 American Cyanamid Injection Facility Hydrogeologic Cross Section
(Ehrlichetal., 1979).
;.•'.' . SANO-ANCW3HAVEL AQUIFER
. . (Fr«th-wanr tupplv)- . • •
; PENSACOLA CLAY ^r^nj^rLrL:
— (Confining b«d)
| [UPPER LIMESTONE FLORIDAN AQUIFER |
—-T_r_r_nrL7LT^ir^. BUCATUNNACLAY —_^-_^-_^
^j^^j^j-^j-j^-— (Confining b*d) I ~—
I I I LOWER LIMESTONE.FLORIDAN, AQUIFER I
, 1/1 r. ,-. ,-, .-,
part of the Floridan aquifer above the confining layer.
A downgradient, deep monitoring well was placed in
the injection zone 1,000 ft southeast of the injection
well (see Figure 7-7). Another shallow well, located
2 miles southeast of the injection site at the Univer-
sity of Florida's Everglades Experiment Station, has
also been monitored for near-surface effects.
Acetate ions from the injected waste were detected in
the deep monitoring well 1,000 ft southeast of the injec-
tion well in early 1967, a matter of months after injec-
tion began (Garcia-Bengochea and Vernon, 1970). In
1971, about 27 months after injection began, evidence
of waste migration was detected at a shallow monitor-
ing well in the upper part of the Floridan aquifer (see
Section 7.4.3 for discussion of geochemical evidence).
Dissolution of the carbonate confining layer by the
acidic waste was the main reason for the upward
migration. However, the lower density of the injected
wastes compared with that of the formation waters
(0.98 g/mL vs. 1,003 g/mL) served to accelerate the
rate of upward migration (Kaufman et al., 1973). In
an attempt to prevent further upward migration, the
injection well was deepened to 2,242 ft, and the inner
casing was extended and cemented to 1,938 ft. When
waste injection was resumed, evidence of upward
migration to the shallow aquifer was observed only 15
months later. By late 1973, 7 years after injection
began, the waste front was estimated to have migrated
0.6 to 1 mile from the injection well (Kaufman and Mc-
Kenzie, 1975).
The injection well was deepened a third time, to a depth
of 3,000 ft (McKenzie, 1976). A new, thicker confining
zone of dense carbonate rock separates the current injec-
tion zone from the previous zone (see Figure 7-8^he
current injection zone is not shown). As of early 1989, the
wastes were still contained in the deepest injection zone
(Haberfeld, 1989b).
150
-------
7.4.2 Injectlon/Confinlng-Zone Lithologyand
Chemistry
The wastes are injected into the lower part of the
carbonate Floridan aquifer, which is extremely per-
meable and cavernous (see Figure 7-8). The natural
direction of ground-water flow is to the southeast
(see Figure 7-7). The confining layer is 150 ft of
dense carbonate rocks. The chloride concentration in
the upper part of the injection zone is 1,650 mg/L, in-
creasing to 15,800 mg/L near the bottom of the for-
mation (Kaufman et al., 1973). The sources used for
this case study did not provide any data on the cur-
rent injection zone. The native fluid was basically a
sodium-chloride solution but also included significant
quantities of sulfate (1,500 mg/L), magnesium (625
mg/L), and calcium (477 mg/L). See Table 6-6 in
Chapter Six for additional data on the chemistry of
the formation water.
7.4.3 Chemical Processes Observed
Neutralization of the injected acids by the limestone for-
mation led to concentrations of calcium, magnesium,
and silica in the waste solution that were higher than
those in the unneutralized wastes. Anaerobic decom-
position of the organic matter in the injected waste
apparently occurred through the action of both
sulfate-reducing and methanogenic bacteria. Sulfate-
reducing bacteria were observed in the injected wastes
that were allowed to backflow to the surface. Sulfate
levels in the native ground water declined by 45%, and
the concentration of hydrogen sulfide increased by
1,600%. Methane fermentation (reduction of CC-2 to
CH4) was also inferred from the presence of both
gases in the backflow fluid, but the presence of
methanogenic bacteria was not confirmed. Increased
hydrogen sulfide concentrations produced by the bac-
teria during biodegradation and the subsequent
decrease in sulfate/chloride ratio in the observation
wells were taken as indicators of upward and lateral
migration. Migration into the shallow monitoring well
was also indicated by a decline in pH from around 7.8
to 6.5, caused by mixing with the acidic wastes.
Chemical analyses of the backf lowed injected waste that
had been in the aquifer for about 2.5 months (for which
some dilution had occurred) indicated that chemical
oxygen demand (COD) was about half that of the
Figure 7-7 Index Map of Belle Glade Area and Potentiometric-Surface Map of the Floridan
Aquifer in South Florida (Kaufman et al., 1973).
EXPLANATION
50
POTENTIOMETRIC CONTOUR. FLORIDA*
COUIFER. JULY 6-17, 1961. DATUM IS
MEAN SEA LEVEL
AREA OF ARTESIAN FLOW
FLORIDAN AQUIFER
SHALLOW
MONITOR
WELL
QftOWERS —
QUAKER OATS
PLANTS
INDICATES DIRECTION OF GROUND WATER
MOVEMENT « | ? MILES
0 IO 2O 30 40 5O MILES
i i i l I I
151
-------
Figure 7-8 Generalized Hydrogeologic Section
between Belle Glade and the Straits of Florida
(Kaufman and McKenzie, 1975).
«- EAST
-40 mi (64 km) —
///7/Dense carbonate rocks ///
1000-
2000-
2500 -
Chlonde concenlcation. Shallow aquifer —
sandl shel1' and """"tone
** , ' '. ' j * °nu nmesio
W777777777777/
Confining beds—dense marl /
Upper part of the
Flondan aquifer—
„., . . permeable carbonat<
Chloride concentration. rock-
1000 mg/L
///// Dense carbonate rocks ///
Chloride concentration.
1650 mg/L
.
Lower part of the
Floridan aquifer —
permeable carbonate
rocks
Chloride concentration.
7000 mg/L
Highly permeable
cavernous
1l5,800 mg/L carbonate rocks
77 7/7/7/77 7/7 7777
Dense carbonate rocks
original waste. Samples that had been in residence for
about 5 months had a COD approximately one-quarter
that of the original waste (12,200 mg/L in the original
waste compared with 4,166 mg/L in the samples). The
percent reduction in COD resulting from bacterial ac-
tion rather than dilution was not estimated.
7.5 Case Study No. 4: Wilmington,
North Carolina
7.5.1 Injection-Facility Overview
The Hercules Chemical, Inc. (now Hercufina, Inc.),
facility, 4 miles north of Wilmington, North Carolina, at-
tempted deep-well injection of its hazardous wastes
from May 1968 to December 1972, but had to discon-
tinue injection because of waste-reservoir incom-
patibility and unfavorable hydrogeologic conditions.
The U.S. Geological Survey conducted extensive
geochemical studies of this site until the well was aban-
doned (Leenheer and Malcolm, 1973; Peek and Heath,
1973; Leenheer et al., 1976a,b). Biodegradation
processes were also studied (DiTommaso and Elkan,
1973; Elkan and Horvath, 1977). More geochemical-
fate processes affecting injected organic wastes have
been documented at this site than at any other.
Hercules Chemical produced an acidic organic waste
derived from the manufacture of dimethyl terphthalate,
which is used in the production of synthetic fiber. The
average dissolved organic carbon concentration was
about 7,100 mg/L and included acetic acid, formic acid,
p-toluic acid, formaldehyde, methanol, terphthalic acid,
and benzoic acid. The pH ranged from 3.5 to 4.0. The
waste also contained traces (less than 0.5 mg/L) of 11
other organic compounds, including dimethyl phthalate, a
listed hazardous waste (see Table 4-8 in Chapter Four).
From May 1968 to December 1972, the waste was in-
jected at a rate of about 300,000 gallons per day. The
first injection well (I-6) was completed to a depth of 850
to 1,025 ft (i.e., cased from the surface to 850 ft with
screens placed in the most permeable sections of the
injection zone to a depth of 1,025 ft). One shallow ob-
servation well (No. 3) was placed 50 ft east of the injec-
tion site at a depth of 690 ft. Four deep monitoring
wells (Nos. 1, 2,4, and 5) were also placed in the injec-
tion zone, one at 50 ft and three at 150 ft from the
injection well (see Figure 7-9).
The injection well became plugged after a few months
of operation because of the reactive nature of the
wastes and the tow permeability of the injection zone.
The actual plugging process was caused both by
reprecipitation of the initially dissolved minerals and by
plugging of pores by such gaseous products as carbon
dioxide and methane. When the first well failed, a
second injection well (I-7A) was drilled into the same
injection zone about 5,000 ft north of the first, and injec-
tion began in May 1971. Nine additional monitoring
wells (three shallow, Nos. 8, 9, and 13, and six deep,
Nos. 7, 11, 12, 14, 15, and 16) were placed at
distances ranging from 1,500 to 3,000 ft from the
second injection well (see Figure 7-9). Injection was
discontinued in 1972 after the operators determined
that the problems of low permeability and waste-
reservoir incompatibility could not be overcome.
Monitoring of the waste movement and subsurface en-
vironment continued into the mid-1970s in the three
monitoring wells located 1,500 to 2,000 ft from the
injection wells.
Within 4 months, the waste front had passed the deep
observation wells located within 150 ft of the injection
well (Nos. 1, 2, 4, and 5). About 9 months after injec-
tion began, leakage into the aquifer above the confining
layer was observed (Well No. 3). This leakage was ap-
parently caused by the increased pressures created by
formation plugging and by the dissolution of the confin-
ing beds and the cement grout surrounding the well
casing of several of the deep monitoring wells, caused
by organic acids.
Eight months after injection began in the second injec-
tion well, wastes had leaked upward into the adjacent
shallow monitoring well (Well No. 9). The leak ap-
parently was caused by the dissolution of the cement
grout around the casing. In June 1972, 13 months after
152
-------
Figure 7-9 Map of Wilmington, North Carolina, Waste-Injection and Observation Wells
(Leenheerand Malcolm, 1973).
IS
A
WILMINGTON
4 MILES
14
PLANT
II
A
I-7A
NORTH
I \ 16
/ x *
V \
2
A
I
A
1-6 3
• A
13
A
SCALE
• -INJECTION WELL
A-OBSERVATION WELL
PLANT LOCATION - 34° 19' N 077° se'
injection began in the second well, the waste front
reached the deep monitoring well located 1,500 ft
northwest of the injection well (No. 14), and in August
1972 waste was detected in Well No. 11 (about 1,000 ft
north of injection well I-7A). Waste injection ended in
December 1972. As of 1977, the wastes were treated in
a surface facility (Elkan and Horvath, 1977).
7.5.2 Injection/Confining-Zone Llthologyand
Chemistry
The injection zone consisted of multiple Upper
Cretaceous strata of sand, silty sand, clay, and some
thin beds of limestone (see Figure 7-10). The clay con-
fining layer was about 100 ft thick. The total-dissolved-
solids concentration in the injection-zone formation water
was 20,800 mg/L, with sodium chloride the most abun-
dant constituent (see Table 6-6 in Chapter Six for addi-
tional data on the chemistry of the formation water).
7.5.3 Chemical Processes Observed
A number of chemical processes were observed at
the site (Leenheer and Malcolm, 1976a,b):
• The waste organic acids dissolved carbonate
minerals, alumino-silicate minerals, and iron/
manganese-oxide coatings on the primary minerals
in the injection zone.
• The waste organic acids dissolved and formed
complexes with iron and manganese oxides.
These dissolved complexes reprecipitated when
the pH increased to 5.5 or 6.0 because of
neutralization of the waste by the aquifer
carbonates and oxides.
• The aquifer mineral constituents adsorbed most
waste organic compounds except formaldehyde.
Adsorption of all organic acids except phthalic
acid increased with a decrease in waste pH.
153
-------
Figure 7-10 Diagram Showing Construction
Features and Lithologic Log of
Well 14, Wilmington, North Carolina
(Leenheer and Malcolm, 1973).
PRESSURE
GAUGE
'•'.•'••'.'•-.•'•'.'..Cl_ = 2400-Z800 iiig/L;..'-.v.V-.
•'.•:•.'.'.'.'.'.Cl =9.000-10,000 rng/L.- :'.V..".
y
4 10 STEEL
PIPE CASING
SAND 8 GRAVEL EH CLAY
LIMESTONE f77] CRYSTALLINE ROCK
• Phthalic acid was complexed with dissolved iron.
The concentration of this complex decreased
as the pH increased because the complex
coprecipitated with the iron oxide.
• Biochemical waste transformation occurred at low
waste concentrations, resulting in the production
of methane. Additional microbial degradation of
the waste resulted in the reduction of sulfates to
sulfides and ferric ions to ferrous ions.
When the dilute waste front reached Well No. 14, in
June 1972, microbial populations rapidly increased in
this well, with methanogenesis being the major
degradative process (DiTommaso and Elkan, 1973).
Elkan and Horvath (1977) found greater numbers and
species diversity of microorganisms in Well No. 11,
which contained dilute wastes, than in Well No. 7,
which was uncontaminated. In laboratory experiments,
however, DiTommaso and Elkan (1973) found that
bacterial growth was inhibited as the concentration of
waste increased and could not decompose the waste
at the rate it was being injected.
This case study illustrates the importance of
dissolution/precipitation reactions in determining waste-
reservoir compatibility. Adsorption was observed to
immobilize most of the organic constituents in the
waste except for formaldehyde. As with the
Monsanto case study, biodegradation was an im-
portant process when wastes were diluted by forma-
tion waters, but the process became inhibited when
undiluted waste reached a given location in the
injection zone.
7.6 Case Study No. 5: Illinois
Hydrochloric Acid-Injection Well
7.6.11njection-Facility Overview
This case study is an example of a well blowout
resulting from the neutralization of acid by carbonate
rock. Kamath and Salazar (1986) and Panagiotopoulos
and Reid (1986) both discuss the same incident. Al-
though they do not specify the location, Brower et al.
(1989) identify the site as the Cabot Corporation injec-
tion well, near Tuscola, Illinois.
The waste hydrochloric acid (HCI) injected at the site
was a byproduct of a combustion process at 2,972°F.
When not recovered, the acidic stream was dumped
into holding ponds where it was cooled to about 75°F
before injection. The concentration of injected acid typi-
cally varied from 0.5 to 5% HCI, but ranged as high as
about 30%. (The pH of injected acid that backflowed
during one blowout incident ranged from 0.5 to 1.3.)
The injection well was cased to a depth of about
4,900 ft and extended into dolomite to a total depth
of 5,300 ft. Injection began in the early 1960s and
averaged around 90 gallons per minute (gpm). The
natural fluid level was 200 ft below the wellhead, and
wastes were injected using gravity flow, i.e., the
pressure head of the well when filled to the surface
with fluid was sufficient to inject fluids without pump-
ing under pressure (Kamath and Salazar, 1986).
Between 1973 and 1975, several blowouts caused sur-
face water pollution and fishkills. The most serious oc-
curred in 1975 after unusually high concentrations of
HCI (around 30%) were injected intermittently for
several weeks. The well refused to accept additional
acid under gravity flow. At first the operators thought
the well bore had become plugged, and they pumped a
concentrated calcium-chloride solution down the hole
to dissolve precipitates that might have formed. Shortly
thereafter the well tubing broke, pressure suddenly
154
-------
rose to 450 psi, and a section of the upper tubing
was ejected through the wellhead along with acid
and annulus fluids. Backflow was stopped for a while
by draining cold water from a fire hydrant into the
well at 50 gpm. The well erupted again the next day,
however, with a 10-ft gusher discharging at 250 gpm.
The blowout was brought under control 2 days later
when a blowout preventer was installed.
7.6.2 Injection/Confining-Zone Lithology and
Chemistry
The injection zone was a cavernous dolomite, and
the native ground water was very saline, with TDS
levels ranging from 21,000 to 26,000 mg/L. No infor-
mation was provided on the confining layer, but it is
discussed in Brower et al. (1989) in detail.
7.6.3 Chemical Processes Observed
The HCI dissolved the dolomite, forming carbon dioxide
(CO2) gas. Under normal circumstances this gas
remains in solution, but if the temperature of the acid
and/or the acid concentration exceed certain limits,
COa evolves as a gas and accumulates in the upper
portion of the cavity. The escape of even small
amounts of COa into the injection pipe can serve as a
driving force to reverse the flow of the injected liquids,
because as the COa rises, pressure decreases and the
gas expands.
There is some disagreement as to which parameter is
most critical to gas blowout. Based on analysis of COa
phase behavior at different temperatures and pres-
sures, Kamath and Salazar (1986) concluded that gas
blowout becomes hazardous if the temperature of the
injected HCI exceeds 88°F. Panagbtopoulos and Reid
(1986) concluded that HCI concentration is the critical
factor and that HCI concentrations exceeding 6% will
evolve COa gas and create a blowout hazard. Both
sets of investigators explained the circumstances of
this case study in terms of their respective models.
7.7 Case Study No. 6: Texas
Petrochemical Plant
7.7.1lnjection-Facility Overview
This case study involves an unnamed petrochemical
plant located about 15 miles inland from the Texas
Gulf Coast, described by Donaldson and Johansen
(1973). It illustrates two approaches to injecting in-
compatible waste streams to prevent well plugging
by precipitation: surface treatment and multiple injec-
tion wells.
The plant began full-scale operation in 1962 and
produced acetic, adipic, and propionic acids;
acetaldehyde; butanol; hexamethyldiamine; vinyl
acetate; nylon; and other chemical products from
petroleum-base stocks. The effluent was collected at
waste treatment facilities as two separate mixtures.
Because mixing two waste streams produced consid-
erable precipitation, the waste streams were
processed and injected separately into two wells.
Organic constituents in the first waste stream totaled
about 14,000 mg/L (acetaldehyde, acetaldol, acetic
acid, butanol-1, butyraldehyde, chloroacetaldehyde,
crotonaldehyde, phenol, and propionic acid) and about
5,200 mg/L inorganic constituents. The pH ranged from
4 to 6, and TDS ranged from 3,000 to 10,000 mg/L.
The second waste stream contained amines and
nitrates generated from the manufacture of nylon,
hydrocarbon solvents used in processing, and other
minor constituents. Organic constituents (amyl al-
cohol, cyclohexane, dodecane, hexanol, 1-hexyl-
amine, 1,6-hexylamine, methanol, and valeric acid)
totaled about 4,700 mg/L. Inorganic constituents in
the second waste stream totaled about 21,350 mg/L,
including 7,500 mg/L nitrate and 4,600 mg/L nitrite.
The second waste stream was basic, with a pH from 8
to 10. The composition of the wastes changed over
time when processes changed or a new unit was in-
stalled. Several new process wastes (unspecified) that
were incompatible with either waste stream were made
compatible by adjusting the pH and diluting them.
Injection began in both wells in mid-1963. The injec-
tion zone for Well No. 1 was 45 ft thick beginning at
about 3,400 ft below the surface. Well No. 2 was lo-
cated 2,700 ft north of Well No. 1, and the injection
zone was located between 3,520 and 3,550 ft.
Donaldson and Johansen (1973) mention no
monitoring wells at the site. About 6 years after
injection began, pressure interference from the two
injection wells was observed. During the same
period, the fluid front from Well No. 1 was about 730
ft from the well bore.
7.7.2 Injection/Confining-Zone Lithology and
Chemistry
The injection formation was a loosely consolidated,
fine-grained Miocene sand (see Figures 7-11 and
7-12). The confining strata between the base of the
freshwater aquifer and the injection zone included
about 1,200 ft of relatively impermeable shale and
clay beds with individual zone thickness ranging from
10 to 245 ft.
7.7.3 Chemical Processes Observed
Wellhead pressures increased when injection was
stopped at Well No. 1 for more than 24 hours, apparently
155
-------
Figure 7-11. North-South Cross Section
Showing Oil Wells and Inclination
of Major Formations, Texas
Petrochemical Plant (Donaldson and
Johansen, 1973).
Figure 7-12. Texas Petrochemical Plant
Injection Well (Donaldson and
Johansen, 1973).
Woste injection wells Rwiimnn,
500
1,000
1,500
| 2,000
JE 2,500
(L.
3,000
3,500
4,000
4,500
5,000
-
»s
~~
X
\
\
— ^.^
-~~
._.
X
\
-Ji
^
;--.
•^>v
K- ___
•s ^___
- ^^
N
k-^--,
••• —
— *^
-^,
\
\
- -^_N^
0 5 IO 2
Kilometers
0
i!oy_forrnatiori"
' —
Lissie
-^Formation —
undifferentiated
"* Injection lont
Anahuoc
Formation
Frio
Formation
^
caused by a combination of precipitation reactions
and backflow of sand. Injecting a slug of brine after
every period of interrupted flow solved this
problem. Movement of the main organic con-
stituents (n-hexylamine, butanal, butanol, and
phenol) was assumed to be slowed by adsorption.
This conclusion was based on laboratory adsorption
experiments by involving a different geologic forma-
tion (Cottage Grove sandstone); no direct observa-
tions were made of the injected waste. Section
5.2.2.1 and Table 5-4 in Chapter Five give additional
information on the results of these adsorption experi-
ments.
References*
Barraclough, J. T. 1966. Waste Injection into a Deep
Limestone in Northwestern Florida. Ground Water
4(1):22-24.
Brower, R. D., A. P. Visocky, I. G. Krapac, B. R.
Hensel, G. R. Peyton, J. S. Neaton, and M. Guthrie.
'References with more than six authors are cited with
"et al."
IOO •
200 •
300 •
400 •
t 500 •
i
600
700
800
900
IjOOO
I,IOO
TO M51.2
38 cm hole
27-cm casing
8-cm casing
^- 11-cm tubing
Cement
Sand with clay
Sand with
inoK breaks
"I Waste injection zone
Mine-groin, soft sand
Shale
Sand
State wM sand
1989. Evaluation of Underground Injection of Industrial
Waste in Illinois, Final Report. Illinois Scientific Surveys
Joint Report 2. Illinois State Geological Survey, Cham-
paign, Illinois.
Dean, B. T. 1965. The Design and Operation of a Deep-
Well Disposal System. Water Poll. Control Fed. J.
37:245-254.
DiTommaso, A., and G. H. Elkan. 1973. Role of Bacteria
in Decomposition of Injected Liquid Waste at Wil-
mington, North Carolina. In Symposium on Underground
Waste Management and Artificial Recharge,
J. Braunstein, ed. Pub. No. 110, Int. Assn. of Hydrotogi-
cal Sciences, pp. 585-599.
156
-------
Donaldson, E. C. 1964. Subsurface Disposal of In-
dustrial Wastes in the United States. U.S. Bureau of
Mines Information Circular 8212.
Donaldson, E. C., and R. T. Johansen. 1973. History of a
Two-Well Industrial-Waste Disposal System. In Sym-
posium on Underground Waste Management and Artifi-
cial Recharge, J. Braunstein, ed. Pub. No. 110, Int.
Assn. of Hydrological Sciences, pp. 603-621.
Donaldson, E. C., R. D. Thomas, and K. H.
Johnston. 1974. Subsurface Waste Injection in the
United States, Fifteen Case Histories. U.S. Bureau of
Mines Information Circular 8636.
Ehrlich, G. G., E. M. Godsy, C. A. Pascale, and J.
Vecchioli. 1979. Chemical Changes in an Industrial
Waste Liquid during Post-Injection Movement in a
Limestone Aquifer, Pensacola, Florida. Ground
Water 17(6):562-573.
Elkan, G., and E. Horvath. 1977. The Role of Microor-
ganisms in the Decomposition of Deep Well Injected
Liquid Industrial Wastes. NSF/RA-770102, NTIS PB
268646.
Faulkner, G. L, and C. A. Pascale. 1975. Monitoring
Regional Effects of High Pressure Injection of In-
dustrial Waste Water in a Limestone Aquifer. Ground
Water 13(2) :197-208.
Foster, J. B., and D. A. Goolsby. 1972. Construction of
Waste-Injection Monitoring Wells near Pensacola, Florida.
Florida Bureau of Geology Information Circular 74.
Garcia-Bengochea, J. I., and R. O. Vernon. 1970.
Deep-Well Disposal of Waste Waters in Saline
Aquifers of South Florida. Water Resources Re-
search 6:1464-70.
Goolsby, D. A. 1971. Hydrogeochemical Effects of
Injecting Wastes into a Limestone Aquifer near Pen-
sacola, Florida. Ground Wafer9(1):13-17.
Goolsby, D. A. 1972. Geochemical Effects and
Movement of Injected Industrial Waste in a Lime-
stone Aquifer. In Symposium on Underground Waste
Management and Environmental Implications, Hous-
ton, Texas, T. D. Cook, ed. Am. Assn. Petr. Geol.
Mem. 18, pp. 355-368.
Haberfeld, J. 1989a. Personal communication. Bureau of
Groundwater Protection, Florida Department of Environ-
ment Regulations, Tallahassee, Florida, March 31.
Haberfeld, J. 1989b. Personal communication. Bureau of
Groundwater Protection, Florida Department of Environ-
ment Regulations, Tallahassee, Florida, April 3.
Kamath K., and M. Salazar. 1986. The Role of the Criti-
cal Temperature of Carbon Dioxide on the Behavior of
Wells Injecting Hydrochloric Acid into Carbonate Forma-
tions. In Proc. of the Int. Symp. on Subsurface Injection
of Liquid Wastes, New Orleans. National Water Well As-
sociation, Dublin, Ohio, pp. 638-655.
Kaufman, M. I., and D. J. McKenzie. 1975. Upward
Migration of Deep-Well Waste Injection Fluids in
Floridan Aquifer, South Florida. J. Res. U. S. Geol.
Survey 3(3) :261 -271.
Kaufman, M. I., D. A. Goolsby, and G. L. Faulkner.
1973. Injection of Acidic Industrial Waste into a Saline
Carbonate Aquifer: Geochemical Aspects. In Sym-
posium on Underground Waste Management and Ar-
tificial Recharge, J. Braunstein, ed. Pub. No. 110, Int.
Assn. of Hydrological Sciences, pp. 526-555.
Keely, J. F. 1982. Chemical Time-Series Sampling.
Ground Water Monitoring Review Fall:29-38.
Keely, J. F., and F. Wolf. 1983. Field Applications of
Chemical Time-Series Sampling. Ground Water
Monitoring Review Fall:26-33.
Leenheer J. A., and R. L. Malcolm. 1973. Case His-
tory of Subsurface Waste Injection of an Industrial
Organic Waste. In Symposium on Underground
Waste Management and Artificial Recharge, J.
Braunstein, ed. Pub. No. 110, Int. Assn. of Hydrologi-
cal Sciences, pp. 565-584.
Leenheer, J. A., R. L. Malcolm, and W. R. White.
1976a. Physical, Chemical and Biological Aspects of
Subsurface Organic Waste Injection near Wil-
mington, North Carolina. U.S. Geological Survey
Professional Paper 987.
Leenheer, J. A., R. L. Malcolm, and W. R. White.
1976b. Investigation of the Reactivity and Fate of Cer-
tain Organic Compounds of an Industrial Waste After
Deep-Well Injection. Environ. Sci. Tech. 10(5):445-451.
McKenzie, D. J. 1976. Injection of Acidic Industrial
Waste into the Floridan Aquifer near Belle Glade,
Florida: Upward Migration and Geochemical Interac-
tions. U.S. Geological Survey Open File Report 76-626.
Pascale C. A., and J.B. Martin. 1978. Hydrologic
Monitoring of a Deep-Well Waste-Injection System
near Pensacola, Florida, March 1970-March 1977.
157
-------
U.S. Geological Survey Water Resource Investiga-
tion 78-27.
Peek, H. M., and R. C. Heath. 1973. Feasibility Study
of Liquid-Waste Injection into Aquifers Containing Salt
Water, Wilmington, North Carolina. In Symposium on
Underground Waste Management and Artificial
Recharge, J. Braunstein, ed. Pub. No. 110, Int. Assn.
of Hydrofogical Sciences, pp. 851-875.
Panagiotopoulos, A. Z., and R. C. Reid. 1986. Deep-
Well Injection of Aqueous Hydrochloric Acid. In Proc.
of the Int. Symp. on Subsurface Injection of Liquid
Wastes, New Orleans. National Water Well Associa-
tion, Dublin, Ohio, pp. 610-637.
Reeder, L. R., et al. 1977. Review and Assessment
of Deep-Well Injection of Hazardous Wastes, Vol. IV,
Appendix E. EPA 600/2-77-029d, NTIS PB 269 004.
Vecchioli, J., G. G. Ehrlich, E. M. Godsy, and C. A.
Pascale. 1984. Alterations in the Chemistry of an In-
dustrial Waste Liquid Injected into Limestone near
Pensacola, Florida. In Hydrogeology of Karstic Ter-
rains, Case Histories, Vol. 1, G. Castany, E. Groba,
and E. Romijn, eds. Int. Assn. of Hydrogeologists,
pp. 217-221.
Willis, C. J., G. H. Elkan, E. Horvath, and K. R. Dail.
1975. Bacterial Flora of Saline Aquifers. Ground
Water 13(5) :406-409.
158
-------
APPENDIX A
SECTION AND TABLE INDEX FOR EPA PRIORITY POLLUTANTS
-------
-------
Appendix A. Section and Table Index for EPA Priority Pollutants
Compound
Acenaphtheneb
Acenaphthylene
Acroleinb
Acrylonitrileb
Aldrinb
Anthracene6
Benzeneb
Benzidine
Benzo(a)anthraceneb
Benzo(a)pyreneb
Benzo(b)fluoranthene
Benzo(ghi)perylene
Benzo(k)fluoranthene
Bromodichloromethaneb
4-Bromodiphenyl ether
Bromoform (see tribromomethane)b
Bromomethane (methyl bromide)
Carbon tetrachloride (see tetrachloromethane)b
Chlordane
Chlorodibromomethane (see dibromochloromethane)
4-Chlorodiphenyl ether
Chlorobenzeneb
Chloroethane (ethyl chloride)b
Chloroethene (vinyl chloride)b
bis(2-Chloroethoxy) methane
bis(2-Chloroethyl) ether6
2-Chloroethyl vinyl ether
Chloroform (see trichloromethane)b
bis(2-Chloroisopropyl) ether1*
p-Chloro-m-cresol
Chloromethane (methyl chloride)
bis(Chloromethyl) ether
2-Chlorophenor
Chryseneb
ODD
DDEb
DDT"
Dibenzo(a,h)anthracene
Dibromochloromethane (Chlorodibromomethane)
1 ,2-Dichlorobenzene (o-dichlorobenzene)b
1 ,3-Dichtorobenzene (m-dichlorobenzene)b
Group8
PAH
PAH
P
N
P
PAH
MA
N
PAH
PAH
PAH
PAH
PAH
HAH
HE
HAH
HAH
HAH
P
HAH
HE
MA
HAH
HAH
HE
HE
HE
HAH
HE
MA
HAH
HE
MA
PAH
P
P
P
PAH
HAH
MA
MA
Section
4.3.5
4.3.5
4.3.7
4.3.6
4.3.7
4.3.5
4.3.3
4.3.6
4.3.5
4.3.5
4.3.5
4.3.5
4.3.5
4.3.1
4.3.2
4.3.1
4.3.1
4.3.1
4.3.7
4.3.1
4.3.2
4.3.3
4.3.1
4.3.1
4.3.2
4.3.2
4.3.2
4.3.1
4.3.2
4.3.3
4.3.1
4.3.2
4.3.3
4.3.5
4.3.7
4.3.7
4.3.7
4.3.5
4.3.1
4.3.3
4.3.3
Table
4-9
4-9
4-11
4-10
4-11
4-9
4-7
4-10
4-9
4-9
4-9
4-9
4-9
4-5
4-6
4-5
4-5
4-5
4-11
4-5
4-6
4-7
4-5
4-5
4-6
4-6
4-6
4-5
4-6
4-7
4-5
4-6
4-7
4-9
4-11
4-11
4-11
4-9
4-5
4-7
4-7
161
-------
Appendix A (Continued)
Compound
1 ,4-Dichlorobenzene (p-dichlorobenzene)b
3,3'-Dichlorobenzidine
Dichlorodifluoromethane
1 ,1 -Dichloroethane (ethylidene chloride)6
1 ,2-Dichloroethane (ethylene chloride)6
1,1-Dichloroetheneb
trans-1 ,2-Dichloroetheneb
Dichloromethane (see methylene chloride)6
2,4-Dichlorophenolb
1 ,2-Dichloropropaneb
1 ,2-Dichloropropene
Dieldrin6
Diethyl phthalate6
2,4-Dimethyl phenol (2,4-xylenol)
Dimethylnitrosamine6
Dimethyl phthalate6
Di-n-butyl phthalate
4,6-Dinitro-o-cresol
2,4-Dinitrophenolb
2,4-Dinitrotolueneb
2,6-Dinitrotolueneb
Di-n-octyl phthalate
Di-n-propyl nitrosamine
1 ,2-Diphenylhydrazine (hydrazobenzene)
Diphenylnitrosamine
Endosulfan and endosulfan sulfateb
Endrin and endrin aldehyde6
Ethylbenzeneb
Ethyl chloride (see chloroethane)b
Ethylene dichloride (see 1 ,2-dichloroethane)b
bis(2-Ethylhexyi) phthalateb
Ethylidene chloride (see 1 ,1 -dichloroethane)6
Fluoranthene6
Fluorene
Heptachlor
Heptachlor epoxide
Hexachbrobenzene6
Hexachlorobutadiene6
Hexachbrocyclohexane (lindane)6
Hexachlorocyclopentadiene
Group"
MA
N
HAH
HAH
HAH
HAH
HAH
HAH
MA
HAH
HAH
P
PE
MA
N
PE
PE
MA
MA
MA
MA
PE
N
N
N
P
P
MA
HAH
HAH
PE
HAH
PAH
PAH
P
P
MA
HAH
P
HAH
Section
4.3.3
4.3.6
4.3.1
4.3.1
4.3.1
4.3.1
4.3.1
4.3.1
4.3.3
4.3.1
4.3.1
4.3.7
4.3.4
4.3.3
4.3.6
4.3.4
4.3.4
4.3.3
4.3.3
4.3.3
4.3.3
4.3.4
4.3.6
4.3.6
4.3.6
4.3.7
4.3.7
4.3.3
4.3.1
4.3.1
4.3.4
4.3.1
4.3.5
4.3.5
4.3.7
4.3.7
4.3.3
4.3.1
4.3.7
4.3.1
Table
4-7
4-10
4-5
4-5
4-5
4-5
4-5
4-5
4-7
4-5
4-5
4-11
4-8
4-7
4-10
4-8
4-8
4-7
4-7
4-7
4-7
4-8
4-10
4-10
4-10
4-11
4-11
4-7
4-5
4-5
4-8
4-5
4-9
4-9
4-11
4-11
4-7
4-5
4-11
4-5
162
-------
Appendix A (Continued)
Compound
Hexachloroethane
lndeno(1 ,2,3-cd)pyrene
Isophorone
Lindane (see hexachlorocyclohexane)b
Methyl bromide (see bromomethane)
Methyl chloride (see chloromethane)b
Methylene chloride (see dichloromethane)b
Methyl chloroform (see 1 ,1 ,1 -trichloroethane)b
Naphthalene6
Nitrobenzeneb
2-Nitrophenolb
4-Nitrophenolb
PCB (see polychlorinated biphenyls)b
Pentachlorophenol6
Perchloroethylene (see tetrachloroethene)b
Phenanthreneb
Phenol6
Polychlorinated biphenyls (PCB)b
Pyreneb
Tetrachlorodibenzodioxin
1,1 ,2,2-Tetrachloroethaneb
Tetrachloroethene(perchloroethylene)b
Tetrachloromethane (carbon tetrachloride)b
Toluene6
Tribromomethane (bromoform)b
1 ,2,4-Trichlorobenzene6
1 ,1 ,1-Trichloroethane (methyl chloroform)6
1,1,2-Trichloroethaneb
Trichlorethene6
Trichloromethane (chloroform)b
Trichlorofluoromethaneb
2,4,6-Trichlorophenolb
Vinyl chloride (see chlorethene)b
Vinylidiene chloride (see 1,1-dichloroethene)b
2,4-Xylenol (see 2,4-Dimethyl phenol)
Group*
HAH
PAH
P
P
HAH
HAH
HAH
HAH
PAH
MA
MA
MA
PAH
MA
HAH
PAH
MA
PAH
PAH
P
HAH
HAH
HAH
MA
HAH
MA
HAH
HAH
HAH
HAH
HAH
MA
HAH
HAH
MA
Section
4.3.1
4.3.5
4.3.7
4.3.7
4.3.1
4.3.1
4.3.1
4.3.1
4.3.5
4.3.3
4.3.3
4.3.3
4.3.5
4.3.3
4.3.1
4.3.5
4.3.3
4.3.5
4.3.5
4.3.7
4.3.1
4.3.1
4.3.1
4.3.3
4.3.1
4.3.3
4.3.1
4.3.1
4.3.1
4.3.1
4.3.1
4.3.3
4.3.1
4.3.1
4.3.3
Table
4-5
4-9
4-11
4-11
4-5
4-5
4-5
4-5
4-9
4-7
4-7
4-7
4-9
4-7
4-5
4-9
4-7
4-9
4-9
4-11
4-5
4-5
4-5
4-7
4-5
4-7
4-5
4-5
4-5
4-5
4-5
4-7
4-5
4-5
4-7
aHAH - halogenated aliphatic hydrocarbon, HE = halogenated ether, MA » monocyclic aromatic, PE = phthalate ester,
PAH » polycyclic aromatic hydrocarbon, N - nitrogenous organic, P - pesticide.
6Others sources of information on partition coefficients and biodegradation for this compound can be found in Appendix B.
163
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-------
APPENDIX B
GROUND-WATER CONTAMINANT FATE (ADSORPTION AND BIODEGRADATION)
AND TRANSPORT STUDIES INDEXED BY ORGANIC COMPOUND
This appendix lists organic compounds for which
data on ground-water retardation/ partition coeffi-
cients and/or biodegradation have been cited in pub-
lished scientific papers and reports. Over 150
organic compounds are listed. Compounds listed in
Appendix A, for which information is available in
other tables and in the text of this reference guide,
are indicated with an asterisk.
A large body of literature is available on adsorption
and biodegradation of pesticides, but only pesticides
that are priority pollutants are included in this appen-
dix. Some review papers that provide data on par-
titioning and biodegradation of other pesticides
include: Hamaker and Thompson (1972), Hamaker
(1972), Crosby (1973), and Rao and Davidson
(1980).
Compounds are listed in alphabetical order. The al-
phabetical location is arranged without the isomer
(cis-, trans-, o-, m-, p-) and substituent-number
designation; isomers and substituent numbers for
compounds with the same chemical compositions
are placed in alphabetical and numerical order.
References listed in this appendix are also coded to
indicate whether they present data based on field
studies (F), laboratory studies (L), and/or whether
quantitative models were developed or tested in the
study (M).
-------
-------
Appendix B Ground-Water Contaminant Fate (Adsorption and Biodegradation) and Transport Studies
Indexed by Organic Compound
Type of Study
Compound
Ground-Water Retardation /
Partition Coefficient
Biodegradation
Acenaphthene*
Aceto nit rile
Acrolein*
Acrylonitrile*
Aldrin*
2-Aminoanthracene
6-Aminochrysene
m-Aminophenol
Aniline
Anisole
Anthracene*
Benzene*
Benzo(a)anthracene*
Benzo(a)pyrene*
Benzoate
Benzonitrile
Biphenyl
Bromobenzene
Bromochloromethane
Karickhoff(1984)
Karickhoff(1984)
Chiouetal. (1983)
Karickhoffetal. (1979)
Nkedi-Kizzaetal. (1985)
Rao etal. (1985)
Schwarzenbach and Westall (1981)
Rogers and McFarlane (1981)
Piet and Smeenk (1985)
Chiouetal. (1977)
Chiouetal. (1983)
Karickhoffetal. (1979)
Rao etal. (1985)
Barber et al. (1988)
Roberts etal. (1980)
Rao etal. (1985)
Chiouetal. (1977)
Rao etal. (1985)
Barber etal. (1988)
Wilson etal. (1985b)
Grula and Grula (1976)
Kobayashi and Rittmann (1982)
Ehrlichetal. (1979)
Kobayashi and Rittmann (1982)
Kobayashi and Rittmann (1982)
Aelionetal. (1987)
Aelionetal. (1987)
Ehrlich etal. (1983)
Kobayashi and Rittmann (1982)
Evans etal. (1965)
Abbot and Gledhill (1971)
Jamison et al. (1971)
Wood etal. (1985)
Kobayashi and Rittmann (1982)
Wilson, J.T., etal. (1986)
Wilson, B.H., etal. (1987)
Barker and Patrick (1985)
Mahadevaiah and Miller (1986)
Kobayashi and Rittmann (1982)
Kobayashi and Rittmann (1982)
Barnsley(1975)
Gibson and Suflita (1986)
Kobayashi and Rittmann (1982)
Kobayashi and Rittmann (1982)
167
-------
Appendix B (Continued)
Compound
Type of Study
Ground-Water Retardation /
Partition Coefficient
Biodegradation
Bromodichloromethane*
Bromoform*
n-Butylbenzene
sec-Butylbenzene
Butyronitrile
Carbon tetrachloride*
Chlorobenzene*
Chlorobenzoate
(various isomers)
Wilson etal. (1981)
Roberts etal. (1985)
Piet and Smeenk (1985)
Roberts etal. (1986)
Curtis etal. (1986)
Roberts etal. (1982)
Schwarzenbach and Westall (1981)
Schwarzenbach and Giger (1985)
Bedientetal. (1983)
Rogers and McFariane (1981)
Chiouetal. (1977)
Curtis etal. (1986)
Roberts etal. (1986)
Roberts etal. (1985)
Roberts etal. (1980)
Schwarzenbach and Westall (1981)
Schwarzenbach and Giger (1985)
Chiou etal. (1977)
Wilson etal. (1981)
Roberts etal. (1982)
Chiou etal. (1983)
Winters and Lee (1987)
Voice etal. (1983)
Voice and Weber (1985)
Rao etal. (1985)
Barber etal. (1988)
Rittmanetal. (1980)
Bouweretal. (1981)
Bouwer and McCarty (1984)
Wilson etal. (1983a)
Wilson etal. (1985a)
Wood etal. (1985)
Bouwer and McCarty (1983b)
Kobayashi and Rittmann (1982)
Roberts etal. (1982)
Rittmann et al. (1980)
Bouwer and McCarty (1984)
Wood etal. (1985)
Bouwer and McCarty (1983b)
Kobayashi and Rittmann (1982)
Roberts etal. (1982)
Roberts etal. (1986)
Grula and Grula (1976)
Bouwer and McCarty (1984)
Wood etal. (1985)
Bouwer and McCarty (1983a)
Bouwer and McCarty (1983b)
Kobayashi and Rittmann (1982)
Parsons etal. (1985)
Rittmann et al. (1980)
Bouwer and McCarty (1984)
Wilson etal. (1985a)
Wood etal. (1985)
Wilson etal. (1983a,b)
Bouwer and McCarty (1983b)
Wilson, J.T., etal. (1986)
Kobayashi and Rittmann (1982)
168
-------
Appendix B (Continued)
Type of Study
Compound
Ground-Water Retardation /
Partition Coefficient
Biodegradation
3-Chlorobenzoate
4-Chlorobenzoate
4-Chlorobiphenyl
Chlorocresol
Chlorodibromomethane
Chloroethane*
Chloroethene
(see Vinyl chloride)
bis(2-Chloroethyl)-
ether*
Chloroform*
Winters and Lee (1987)
Piet and Smeenk (1985)
Wilson etal. (1981)
Roberts etal. (1985)
Schwarzenbach and Giger (1985)
Piet and Smeenk (1985)
Chiou etal. (1977)
Wilson etal. (1981)
Roberts etal. (1982)
bis(2-Chloroisopropyl) ether*
Chloromethane
(see Methyl chloride)
p-Chlorophenol
2-ChlorophenoP
3-Chlorophenol
4-Chlorophenol
Chlorophenoxyphenol
Gibson and Suflita (1986)
Gibson and Suflita (1986)
Kobayashi and Rittmann (1982)
Bouwer and McCarty (1984)
Bouwer and McCarty (1983b)
Wood etal. (1985)
Bouwer etal. (1981)
Kobayashi and Rittmann (1982)
Roberts etal. (1982)
Wood etal. (1985)
Kobayashi and Rittmann (1982)
Rittmann etal. (1980)
Bouwer and McCarty (1984)
Wilson etal. (19853)
Wood etal. (1985)
Bouwer etal. (1981)
Wilson etal. (1983b)
Bouwer and McCarty (1983a)
Bouwer and McCarty (1983b)
Kobayashi and Rittmann (1982)
Roberts etal. (1982)
Kobayashi and Rittmann (1982)
Aelionetal. (1987)
Suflita and Miller (1985)
Gibson and Suflita (1986)
Suflita and Miller (1985)
Gibson and Suflita (1986)
Suflita and Miller (1985)
Gibson and Suflita (1986)
Dec and Bollag (1988)
Kobayashi and Rittmann (1982)
Johnson etal. (1985)
169
-------
Appendix B (Continued)
Type of Study
Compound
Ground-Water Retardation /
Partition Coefficient
Biodegradation
Chrysene*
Cresol
(various isomers)
DDE*
DDT*
Dibenzanthracene
Dibenzofuran
Dibromochloromethane
(see Chlorodibromomethane)
1,2-Dibromomethane
Dibutyl phthalate
Dichlorobenzene
1,2-Dichlorobenzene*
1,3-Dichlorobenzene*
1,4-Dichlorobenzene*
Rao etal. (1985)
Tomsonetal. (1985)
Bedientetal. (1983)
Hutchins et al. (1984)
Chiou etal. (1977)
Chiou etal. (1977)
Steinberg etal. (1987)
Chiou etal. (1979)
Tomsonetal. (1985)
Bedient etal. (1983)
Hutchins etal. (1984)
Piet and Smeenk (1985)
Barber etal. (1988)
Bedient etal. (1983)
Chiou etal. (1979)
Hutchins etal. (1984)
Tomsonetal. (1985)
Roberts etal. (1986)
Barber etal. (1988)
Roberts etal. (1980)
Hassettetal. (1980)
Chiou etal. (1979)
Chiou etal. (1983)
Curtis etal. (1986)
Roberts etal. (1980)
Chiou etal. (1983)
Roberts etal. (1980)
Schwarzenbach and Westall (1981)
Schwarzenbach and Giger (1985)
Wilson etal. (1981)
Chiou etal. (1983)
Wu and Gschwend (1986)
Smolenski and Suflita (1987)
Kobayashi and Rittmann (1982)
Goerlitzetal. (1985)
Dobbins and Raender (1987) [m-J
Aelion etal. (1987) [m-]
Kobayashi and Rittmann (1982)
Kobayashi and Rittmann (1982)
Wilson etal. (1985b)
Bouwer and McCarty (1984)
Bouwer and McCarty (1983b)
Wilson etal. (1983a)
Wilson, B.H., etal. (1986)
Aelion etal. (1987)
Roberts etal. (1986)
Wood etal. (1985)
Rittmann etal. (1980)
Bouwer and McCarty (1983b)
Kobayashi and Rittmann (1982)
Kuhnetal. (1985)
Rittmann et al. (1980)
Bouwer and McCarty (1983b)
Rittmann etal. (1980)
Bouwer and McCarty (1984)
Bouwer and McCarty (1983b)
Kobayashi and Rittmann (1982)
170
-------
Appendix B (Continued)
Type of Study
Compound
Ground-Water Retardation /
Partition Coefficient
Biodegradation
2,3-Dichlorobenzene
3,4-Dichlorobenzoate
3,5-Dichlorobenzoate
Dichlorobromomethane
(see Bromodichloromethane)
Dichloroethane
1,1-Dichloroethane*
1,2-Dichloroethane*
Dichloroethene
1,1-Dichlorethene*
cis- and trans-
1,2-Dichloroethene*
1,1-Dichloroethylene
1,2-Dichloroethylene
cis- and trans-
1,2-Dichloroethylene
Dichloromethane
(see Methylene chloride)
2,3-Dichlorophenol
2,4-Dichlorophenol*
2,5-Dichlorphenol
2,6-Dichlorophenol
3,4-Dichlorophenol
1,2-Dichloropropane*
1,3-Dichlorpropylene
Wilson etal. (1981)
Hassettetal. (1980)
Chiou etal. (1979)
Barber etal. (1988)
Schellenberg etal. (1984)
Bedientetal. (1983)
Schellenberg etal. (1984)
Johnson etal. (1985)
Johnson etal. (1985)
Chiou etal. (1979)
Kobayashi and Rittmann (1982)
Gibson and Suflita (1986)
Kobayashi and Rittmann (1982)
Gibson and Suflita (1986)
Kobayashi and Rittmann (1982)
Kobayashi and Rittmann (1982)
Wilson etal. (1985a)
Wood etal. (1985)
Wilson etal. (1983b)
Wilson, B.H., etal. (1986)
Bouwer and McCarty (1984)
Wilson etal. (1985a)
Wood etal. (1985)
Bouwer and McCarty (1983a)
Wood etal. (1985)
Kobayashi and Rittmann (1982)
Wood etal. (1985)
Kobayashi and Rittmann (1982)
Barrio-Lageetal. (1986)
Kobayashi and Rittmann (1982)
Barrio-Lage et al. (1986) [cis]
Kobayashi and Rittmann (1982)
Gibson and Suflita (1986)
Dec and Bollag (1988)
Suflita and Miller (1985)
Gibson and Suflita (1986)
Suflita and Miller (1985)
Gibson and Suflita (1986)
Kobayashi and Rittmann (1982)
171
-------
Appendix B (Continued)
Type of Study
Compound
Ground-Water Retardation /
Partition Coefficient
Biodegradation
Dieldrin*
Diethanolamine
Diethylamine
m-Diethylbenzene
Diethyl phthalate*
Dimethylamine
1,4-Dimethylbenzene
Dimethylnitrosamine*
Dimethyl phthalate*
2,4-Dinitrophenol*
Dinitrotoluene
2,4-Dinitrotoluene*
2,6-Dinitrotoluene*
Diphenyl ether
Endosulfan*
Endrin*
Ethylbenzene*
Ethyl chloride
(see Chloroethane)
Ethylene dibromide
bis(2-Ethylhexyl)
phthalate*
o-Ethyltoluene
Fluoranthene*
Rao etal. (1985)
Tomsonetal. (1985)
Bedientetal. (1983)
Hutchinsetal. (1984)
Piet and Smeenk (1985)
Schwarzenbach and Westall (1981)
Schwarzenbach and Giger (1985)
Tomsonetal. (1985)
Bedient etal. (1983)
Hutchins et al. (1984)
Schwarzenbach et al. (1988)
Piet and Smeenk (1985)
Chiou etal. (1977)
Piet and Smeenk (1985)
Rao etal. (1985)
Chiou etal. (1983)
Rogers and McFarlane (1981)
Tomsonetal. (1985)
Bedient etal. (1983)
Hutchins etal. (1984)
Bedient etal. (1983)
Hutchins et al. (1984)
Tomsonetal. (1985)
Kobayashi and Rittmann (1982)
Boethling and Alexander (1979)
Boethling and Alexander (1979)
Beothling and Alexander (1979)
Kobayashi and Rittmann (1982)
Kobayashi and Rittmann (1982)
Kobayashi and Rittmann (1982)
Kobayashi and Rittmann (1982)
Kobayashi and Rittmann (1982)
Bouwer and McCarty (1984)
Horvath(1972)
Bouwer and McCarty (1983b
Colwell and Sayler (1976)
Kobayashi and Rittmann (1982)
Barnsley(1975)
172
-------
Appendix B (Continued)
Compound
Type of Study
Ground-Water Retardation /
Partition Coefficient
Biodegradation
Fluorene*
Fluorobenzene
Heptaldehyde
Hexachlorobenzene*
Hexachlorobiphenyl
Hexachlorobutadiene*
Hexachlorocyclohexane
(see Lindane)
Hexchloroethane
lodobenzene
Isoquinoline
Lindane*
Methoxychlor
9-Methyl anthracene
3-Methyl benzoate
Methyl chloride*
Methyl chloroform
(see 1,1,1 -Trichloroethane)
Methyl naphthalene
1-Methyl naphthalene
2-Methyl naphthalene
Methylene chloride*
Monochlorophenol
Chiouetal. (1977)
Rao etal. (1985)
Roberts etal. (1980)
Karickhoff(1984)
Karickhoff and Morris (1985)
Karickhoff etal. (1979)
Schwarzenbach and Giger (1985)
Schwarzenbach and Giger (1985)
Curtis etal. (1986)
Roberts etal. (1986)
Chiouetal. (1977)
Rao etal. (1985)
Barber etal. (1988)
Chiouetal. (1985)
Karickhoff etal. (1979)
Karickhoff etal. (1979)
Barber etal. (1988)
Roberts etal. (1980)
Bedientetal. (1983)
Roberts etal. (1980)
Karickhoff etal. (1979)
Bedient etal. (1983)
Hutchinsetal. (1984)
Tomson et al. (1985)
Wilson etal. (1985b)
Rittmanetal. (1980)
Kobayashi and Rittmann (1982)
Griddle etal. (1986)
Pereira and Rostad (1987)
Kobayashi and Rittmann (1982)
Kobayashi and Rittmann (1982)
Kobayashi and Rittmann (1982)
Wood etal. (1985)
Kobayashi and Rittmann (1982)
Wilson etal. (1985b)
Wilson etal. (1985b)
Wood etal. (1985)
Kobayashi and Rittmann (1982)
Kobayashi and Rittmann (1982)
173
-------
Appendix B (Continued)
Type of Study
Compound
Ground-Water Retardation /
Partition Coefficient
Biodegradation
Naphthalene*
Nitrilotriacetic acid
Nitrobenzene*
p-Nitrophenol
2-Nitrophenol*
4-Nitrophenol*
Pentachlorobenzene
Pentachlorophenol*
Perchlorethylene
(see Tetrachloroethene)
Phenanthrene*
Phenol*
Bedientetal. (1983)
Roberts etal. (1980)
Tomsonetal. (1985)
Schwarzenbach and Westall (1981)
Chiouetal. (1977)
Karickhoff etal. (1979)
Winters and Lee (1987)
Voice etal. (1983)
Voice and Weber (1985)
Rao etal. (1985)
Piet and Smeenk (1985)
Barber etal. (1988)
Hutchinsetal. (1984)
Wilson etal. (1981)
Piet and Smeenk (1985)
Schwarzenbach et al. (1988)
Schwarzenbach et al. (1988)
Wu and Gschwend (1986)
Karickhoff and Morris (1985)
O'Connor etal. (1984)
Choi and Aomine (1974a,b)
Schellenberg et al. (1984)
Johnson etal. (1985)
Karickhoff etal. (1979)
Rao etal. (1985)
Scott etal. (1983)
Rittmann et al. (1980)
Bouwer and McCarty (1984)
Ehrlichetal. (1983)
Bouwer and McCarty (1983b)
Wilson etal. (1985b)
Kobayashi and Rittmann (1982)
Ehrlich etal. (1982)
Davies and Evans (1964)
Naumova(1960)
Slavnina(1965)
Ward (1985)
Dunlapetal. (1982)
Kobayashi and Rittmann (1982)
Grula and Grula (1976)
Kobayashi and Rittmann (1982)
Aelion etal. (1987)
Dec and Bollag (1988)
Kobayashi and Rittmann (1982)
Kobayashi ana Rittman (1982)
Evans etal. (1965)
Abbott and Gledhill (1971)
Jamison etal. (1971)
Scott etal. (1983)
Suflita and Miller (1985)
Gibson and Suflita (1986)
Ehrlich etal. (1982)
Ehrlich etal. (1983)
Godsyetal. (1983)
Grula and Grula (1976)
Kobayashi and Rittmann (1982)
Aelion (1987)
174
-------
Appendix B (Continued)
Type of Study
Compound
Ground-Water Retardation /
Partition Coefficient
Biodegradation
Phthalate esters
Polychlorinated
biphenyls (PCBs)*
Propionitrile
Propylbenzene
Pyrene*
Quinoline
Styrene
Tetracene
1,2,3,4-Tetra-
chlorobenzene
1,2,3,5-Tetra-
chlorobenzene
1,1,1-Tetra-
chloroethane
1,1,1,2-Tetra-
chloroethane
1,1,2,2-Tetra-
chloroethane*
Tetrachloroethene*
Weber etal. (1980)
Schwarzenbach and Westall (1981)
Chiou etal. (1977)
Chiou etal. (1983)
Gschwend and Wu (1985)
Rao etal. (1985)
Schwarzenbach and Westall (1981)
Karickhoff etal. (1979)
Karickhoff(1984)
Karickhoff and Morris (1985)
Rao etal. (1985)
Roberts etal. (1980)
Bedientetal. (1983)
Karickhoff etal. (1979)
Wu and Gschwend (1986)
Scwharzenbach and Westall (1981)
Schwarzenbach and Westall (1981)
Schwarzenbach and Giger (1985)
Hassettetal. (1980)
Hassettetal. (1980)
Chiou etal. (1979)
Wilson etal. (1981)
Chiou etal. (1979)
Barber etal. (1988)
Kobayashi and Rittmann (1982)
Colwell and Sayler (1976)
Kobayashi and Rittmann (1982)
Grula and Grula (1976)
Ehrlichetal. (1983)
Kobayashi and Rittmann (1982)
Pereira and Rostad (1987)
Wilson etal. (1983a,b)
Wilson, B.H., etal. (1986)
Kobayashi and Rittmann (1982)
Bouwer and McCarty (1983a)
Bouwer and McCarty (1984)
Parsons etal. (1985)
175
-------
Appendix B (Continued)
Type of Study
Compound
Ground-Water Retardation /
Partition Coefficient
Biodegradation
Tetrachloroethylene
Tetrachloromethane
(see Carbon tetrachloride)
2,3,4,5-Tetra-
chlorophenol
2,3,4,6-Tetra-
chlorophenol
1,2,4,5-Tetra-
methylbenzene
Toluene*
Toxaphene
Tribromomethane
(see Bromoform)
Trichlorobenzene
1,2,3-Trichloro-
benzene
Bedientetal. (1983)
Giger and Molnar-Kubica (1978)
Tomsonetal. 1985
Schwarzenbach and Westall (1981)
Schwarzenbach and Giger (1985)
Chiouetal. (1977)
Wilson etal. (1981)
Curtis etal. (1986)
Roberts etal. (1986)
Hutchinsetal. (1984)
Schellenbergetal. (1984)
Schellenbergetal. (1984)
Johnson etal. (1985)
Schwarzenbach and Westall (1981)
Schwarzenbach and Giger (1985)
Bedientetal. (1983)
Tomsonetal. (1985)
Schwarzenbach and Westall (1981)
Piet and Smeenk (1985)
Chiouetal. (1977)
Wilson etal. (1981)
Hutchins etal. (1984)
Rao etal. (1985)
Bouwer and McCarty (1984)
Wilson etal. (1985a)
Wood etal. (1985)
Bouwer etal. (1981)
Wilson and Wilson (1985)
Wilson etal. (1983a,b)
Bouwer and McCarty (1983a)
Kobayashi and Rittmann (1982)
Kuhnetal. (1985)
Roberts etal. (1982)
Roberts etal. (1986)
Aelionetal. (1987)
Wilson etal. (1985a)
Wilson etal. (1983b)
Kobayashi and Rittmann (1982)
Wilson, J.T., etal. (1986)
Barker and Patrick (1986)
Mahadevaiah and Miller (1986)
Kobayashi and Rittmann (1982)
Kobayashi and Rittmann (1982)
Schwarzenbach and Westall (1981) Kobayashi and Rittmann (1982)
176
-------
Appendix B (Continued)
Compound
Type of Study
Ground-Water Retardation /
Partition Coefficient
Biodegradation
1,2,4-Trichloro-
benzene*
Trichloroethane
1,1,1-Trichloroethane*
1,1,2-Trichloro-
ethane*
Trichloroethene*
Trichloroethylene
Trichlorofluoromethane*
Trichloromethane
(see Chloroform)
2,3,4-Trichlorophenol
2,4,6-Trichlorophenol*
3,4,5-Trichlorophenol
1,2,3-Trimethylbenzene
1,2,4-Trimethylbenzene
1,2,5-Trimethylbenzene
Robertsetal. (1980)
Schwarzenbach and Westall (1981)
Schwarzenbach and Giger (1985)
Wilson etal. (1981)
Chiouetal. (1983)
Wu and Gschwend (1986)
Barber etal. (1988)
Roux and Althoff (1980)
Chiouetal. (1979)
Wilson etal. (1981)
Robertsetal. (1982)
Barber etal. (1988)
Barrio-Lage et al. (1987)
O'Connor et al. (1984)
Schwarzenbach and Giger (1985)
Rogers and McFarlane (1981)
Wilson etal. (1981)
Schellenberg et al. (1984)
Schellenberg et al. (1984)
Johnson etal. (1985)
Schellenberg et al. (1984)
Schwarzenbach and Westall (1981)
Schwarzenbach and Westall (1981)
Schwarzenbach and Giger (1985)
Bouwer and McCarty (1984)
Bouwer and McCarty (1983b)
Kobayashi and Rittmann (1982)
Kobayashi and Rittmann (1982)
Rittmann et al. (1980)
Bouwer and McCarty (1983b)
Bouwer and McCarty (1984)
Wilson etal. (1985a)
Wood etal. (1985)
Wilson et al. (1983b)
Parsons etal. (1985)
Barker etal. (1986)
Vogel and McCarty (1987)
Wilson etal. (1985a)
Wood etal. (1985)
Parsons etal. (1985)
Wilson and Wilson (1985)
Wood etal. (1985)
Bouwer etal. (1981)
Wilson etal. (I983a,b)
Kobayashi and Rittmann (1982)
Wilson, B.H., etal,(1986)
Kobayashi and Rittmann (1982)
Suflita and Miller (1985)
Gibson and Suflita (1986)
Dec and Bollag (1988)
Barker etal. (1986)
177
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Appendix B (Continued)
Compound
Type of Study
Ground-Water Retardation /
Partition Coefficient
Biodegradation
Vinyl chloride*
Vinylidiene chloride
(see1,1-Dichlorethane)
Xylene
(various isomers)
Tomsonetal. (1985)
Bedientetal. (1983)
Hutchinsetal. (1984)
Rao etal. (1985)
Piet and Smeenk (1985)
Wood etal. (1985)
Wilson, J.T.. etal. (1986)
[o-,m-J
Wilson, B. H., etal. (1986) [o-]
Wilson, B.H., etal. (1987)
[m-,p-]
Kuhnetal. (1985) [o-,m-]
Barker etal. (1985) [o-]
Barker and Patrick (1986) [o-,m-]
Mahadevaiah and Miller (1986)
*Priority pollutant; additional data can be found by referring to Appendix A.
178
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