xvEPA
United States
Environmental Protection
Agency
Robert S Kerr Environmental
Research Laboratory
Ada OK 74820
Center for Environmental
Research Information
Cincinnati OH 45268
Technology Transfer
EPA/625/6-87/016
Handbook
Groundwater
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PROPERTY OF THE
OFFICE OF SUPERFUND
EPA/625/6-87/016
March 1987
Handbook
Ground Water
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
U.S. Environmental Protection Agency
Region 5, Library (PL-12J)
77 West Jackson Boulevard, 12th Floor
Chicago, )L 60604-3590
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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 or commercial products does not constitute endorsement or recommendation for use.
This document is not intended to be a guidance or support document for a specific regulatory
program. Guidance documents are available from EPA and must be consulted to address
specific regulatory issues.
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Contents
Chapter Page
PART I FRAMEWORK FOR PROTECTING GROUND-WATER RESOURCES
1 GROUND-WATER CONTAMINATION 1
1.1 Definitions 1
1.2 The Extent of Ground-Water Contamination 1
1.3 General Mechanisms of Ground-Water Contamination 2
1.3.1 Infiltration 2
1.3.2 Direct Migration 3
1.3.3 Interaquifer Exchange 3
1.3.4 Recharge from Surface Water 4
1.4 Sources of Ground-Water Contamination 4
1.5 Movement of Contaminants in Ground Water 7
1.5.1 Contaminant Migration 7
1.5.2 Contaminant Plume Behavior 15
1.6 Summary 17
1.7 References 17
2 GROUND-WATER QUALITY INVESTIGATIONS 21
2.1 Types of Ground-Water Quality Investigations 21
2.1.1 Regional Investigations 21
2.1.2 Local Investigations 21
2.1.3 Site Investigations 21
2.2 Conducting the Investigation 22
2.2.1 Establish the Objectives of the Study 22
2.2.2 Data Collection 22
2.2.3 Field Investigation 23
2.3 Regional Investigations 24
2.4 Local Investigations 26
2.5 Site Investigations 27
2.6 Summary 33
2.7 References 33
3 GROUND-WATER RESTORATION 35
3.1 Subsurface Effects on Contaminant Mobility 35
3.2 Physical Containment Techniques 36
3.2.1 Removal 36
3.2.2 Barriers to Ground-Water Flow 36
3.2.3 Surface Water Controls 37
3.2.4 Limitations of Physical Containment 38
3.3 Hydrodynamic Controls 38
3.3.1 Well Systems 38
3.3.2 Limitations of Hydrodynamic Control 39
3.4 Withdrawal and Treatment 39
3.4.1 Physical 39
in
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Contents (continued)
Chapter Page
3.4.2 Chemical 40
3.4.3 Biological 41
3.4.4 Limitations of Withdrawal and Treatment Techniques 42
3.5 In-Situ Treatment Techniques 43
3.5.1 Chemical/Physical 43
3.5.2 Biodegradation 44
3.6 Treatment Trains 47
3.7 Institutional Limitations on Controlling Ground-Water Pollution 48
3.8 References 49
PART II. SCIENTIFIC AND TECHNICAL BACKGROUND FOR ASSESSING AND
PROTECTING THE QUALITY OF GROUND-WATER RESOURCES
4 BASIC HYDROGEOLOGY 51
4.1 Precipitation 51
4.1.1 Seasonal Variations in Precipitation 51
4.1.2 Types of Precipitation 51
4.1.3 Recording Precipitation 52
4.2 Infiltration 52
4.3 Surface Water 53
4.3.1 Stream Types 54
4.3.2 Stream Discharge Measurements and Records 56
4.4 The Relation Between Surface Water and Ground Water 56
4.4.1 The Regional System 58
4.4.2 Bank Storage 58
4.4.3 Master Depletion Curve 59
4.4.4 Separating a Hydrograph by Graphical Methods 59
4.4.5 Separating a Hydrograph by Chemical Methods 59
4.4.6 Ground-Water Rating Curves 61
4.4.7 Determining Regional Ground-Water Recharge Rates 62
4.4.8 Seepage Measurements 62
4.4.9 Maps of Potential Ground-Water Yield 64
4.4.10 Quality as an Indicator 64
4.4.11 Temperature as an Indicator 69
4.4.12 Flow Duration Curves 69
4.5 Ground Water 73
4.5.1 The Water Table 73
4.5.2 Aquifers and Aquitards 73
4.5.3 Porosity and Hydraulic Conductivity 74
4.5.4 Hydraulic Gradient 74
4.5.5 Potentiometric Surface Map 74
4.5.6 Calculating Ground-Water Flow 76
4.5.7 Interstitial Velocity 78
4.5.8 Transmissivity and Storativity 79
4.5.9 Water-Level Fluctuations 80
4.5.10 Cone of Depression 80
4.5.11 Specific Capacity 80
4.6 References 82
5 MONITORING WELL DESIGN AND CONSTRUCTION 85
5.1 Ground-Water Monitoring Program Goals 85
5.2 Monitoring Well Design Components 86
5.2.1 Location and Number 86
IV
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Contents (continued)
Chapter Page
5.2.2 Diameter 86
5.2.3 Casing and Screen Material 87
5.2.4 Screen Length and Depth of Placement 89
5.2.5 Sealing Materials and Procedures 90
5.2.6 Development 92
5.2.7 Security 93
5.3 Monitoring Well Drilling Methods 95
5.3.1 Geologic Samples 98
5.3.2 Case History 98
5.4 Summary 103
5.5 References 103
6 GROUND-WATER SAMPLING 107
6.1 Introduction 107
6.1.1 Background 107
6.1.2 Information Sources 107
6.1.3 The Subsurface Environment 108
6.1.4 The Sampling Problem and Parameter Selection 108
6.2 Establishing a Sampling Point 110
6.2.1 Well Design and Construction 110
6.2.2 Well Drilling 110
6.2.3 Well Development, Hydraulic Performance and Purging Strategy 111
6.3 Elements of the Sampling Protocol 115
6.3.1 Water-Level Measurement 115
6.3.2 Purging 118
6-3.3 Sample Collection and Handling 118
6.3.4 Quality Assurance/Quality Control 120
6.3.5 Sample Storage and Transport 122
6.4 Summary 123
6.5 References 123
7 GROUND-WATER TRACERS 127
7.1 General Characteristics of Tracers 127
7.2 Public Health Considerations 127
7.3 Direction of Water Movement 127
7.4 Travel Time 127
7.5 Sorption of Tracers and Related Phenomena 128
7.6 Hydrodynamic Dispersion and Molecular Diffusion 128
7.7 Practical Aspects 129
7.7.1 Planning a Test 129
7.7.2 Types of Tracer Tests 130
7.7.3 Design and Construction of Test Wells 134
7.7.4 Injection and Sample Collection 134
7.7.5 Interpretation of Results 135
7.8 Types of Tracers 136
7.8.1 Water Temperature 136
7.8.2 Solid Particles 137
7.8.3 Ions 139
7.8.4 Dyes 139
7.8.5 Some Common Nonionized and Poorly Ionized Compounds 142
7.8.6 Gases 142
7.9 References 145
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Contents (continued)
Chapter Page
8 USE OF MODELS IN MANAGING GROUND-WATER PROTECTION PROGRAMS 149
8.1 The Utility of Models 149
8.1.1 Introduction 149
8.1.2 Management Applications 149
8.1.3 Modeling Contamination Transport 150
8.1.4 Categories of Models 151
8.2 Assumptions, Limitations, and Quality Control 151
8.2.1 Physical Processes 152
8.2.2 Chemical Processes 154
8.2.3 Biological Processes 156
8.2.4 Analytical and Numerical Models 157
8.2.5 Quality Control 158
8.3 Applications in Practical Settings 159
8.3.1 Stereotypical Applications 159
8.3.2 Real-World Applications 160
8.3.3 Practical Concerns 163
8.4 Liabilities, Costs, and Recommendations for Managers 167
8.4.1 Potential Liabilities 167
8.4.2 Economic Considerations 170
8.4.3 Managerial Considerations 180
8.5 References 182
9 BASIC GEOLOGY 185
9.1 Geologic Maps and Cross-Sections 185
9.2 Ground Water in Igneous and Metamorphic Rocks 189
9.3 Ground Water in Sedimentary Rocks 189
9.4 Ground Water in Unconsolidated Sediments 191
9.5 Relationship Between Geology, Climate, and Ground-Water Quality 191
9.6 Minerals 195
9.6.1 Carbonates, Sulfates, and Oxides 195
9.6.2 Rock-Forming Silicates 195
9.6.3 Ores 196
9.7 Rocks 196
9.7.1 Igneous Rocks 196
9.7.2 Metamorphic Rocks 196
9.7.3 Sedimentary Rocks 197
9.8 Weathering 197
9.8.1 Mechanical Weathering 197
9.8.2 Chemical Weathering 197
9.9 Erosion and Deposition 198
9.9.1 Waterborne Deposits 198
9.9.2 Windborne Deposits 198
9.9.3 Glacial Deposits 198
9.10 Geologic Structure 200
9.10.1 Folding 200
9.10.2 Fractures 200
9.11 Geologic Time 201
9.11.1 Rock Units 201
9.11.2 Time and Time-Rock Units 201
9.12 References 202
VI
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Contents (continued)
Chapter Page
APPENDIX SOURCES OF INFORMATION ABOUT GROUND-WATER
CONTAMINATION INVESTIGATIONS 203
SOLID AND HAZARDOUS WASTE AGENCIES 203
U.S. EPA OFFICE OF GROUND-WATER PROTECTION 209
FEDERAL INTERAGENCY GROUND-WATER PROTECTION COMMITTEE 210
VII
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Figures
Number Page
1-1 Growth of the synthetic organic chemical industry in the United States 2
1-2 Plume of leachate migrating from a sanitary landfill on a sandy aquifer
using contours of chloride concentration 4
1-3 Vertical movement of contaminants along an old, abandoned, or improperly
constructed well 4
1 -4 Contaminated water induced to flow from surface water to ground water
by pumping 5
1 -5 Sources of ground-water contamination 7
1-6 Movement of a concentration front by advection only 10
1-7 Movement of a dissolved constituent slug by advection only 10
1 -8 Effect of leakage from a lagoon on a regional flow pattern 10
1 -9 Comparison of advance of contaminant influenced by
hydrodynamic dispersion 11
1-10 Row of contaminated ground water in aquifer with solution porosity 12
1-11 Movement of a concentration front by advection and dispersion 12
1-12 Movement of a dissolved constituent slug by advection and dispersion
as it moves from time period (a) to (b) 12
1-13 Continuous and intermittent sources affected by dispersion 12
1-14 The influence of natural processes on levels of contaminants
downgradient from continuous and slug-release sources 13
1-15 Ion exchange 14
1-16 Metal-ion movement slowed by ion exchange 14
1-17 Benzene and chloride appearance in a monitoring well 16
1-18 Constant release but variable constituent source 17
1-19 Effects of density on migration of contaminants 18
1 -20 Changes in plumes and factors causing the changes 19
2-1 Location of wells with nitrate exceeding 10 mg/l in Region 7 24
2-2 Generalized rock types with high nitrate concentrations in Region 7 25
2-3 Generalized geologic map of a local investigation 27
2-4 Geologic cross section showing downdip change in water quality 28
2-5 Geologic cross section for the site investigation 29
2-6 Map showing thickness of shale overlying the uppermost aquifer 30
2-7 Potentiometric surface of the uppermost aquifer 30
2-8 Relation between precipitation and water level 31
2-9 Relation between precipitation and nitrate concentration 32
4-1 Distribution of annual average precipitation in Oklahoma, 1970-79 52
4-2 Infiltration capacity decreases with time during a rainfall event 53
4-3 Relation between grain size and field capacity and wilting point 53
4-4 Relation between water table and stream type 54
4-5 Water quality data for Cottonwood Creek near Navina, Oklahoma 55
4-6 A generalized stream stage vs discharge rating curve 56
4-7 Stream hydrograph showing definition of terms 56
4-8 Stream discharge record for Cottonwood Creek near Navina, Oklahoma 57
VIII
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Figures (continued)
Number Page
4-9 The chemical quality of ground water commonly changes along a flow
path in the regional system as water flows from areas of recharge to
areas of discharge 58
4-10 Movement of water into and out of bank storage along a stream in Indiana 60
4-11 The shape of ground-water depletion curves changes with the seasons 60
4-12 A stream hydrograph can be separated by three different methods 61
4-13 Schematic showing the contribution of water from different aquifers to
Econfina Creek, Florida 61
4-14 Hydrographs showing the discharge, specific conductance, and computed
ground-water runoff in Four Mile Creek, Iowa 62
4-15 Rating curve of mean ground-water level compared with base flow of
Beaverdam Creek, Maryland 63
4-16 Hydrograph of Brandywine Creek, Chadd's Ford, Pennsylvania, 1952-1953 ... 63
4-17 Rating curve of mean ground-water level and base flow in the Panther Creek
basin, Illinois 64
4-18 Discharge and low flow indices of the Scioto River in central Ohio are strongly
influenced by local geologic conditions 65
4-19 Fish populations are controlled by discharge of mineralized water from an
underlying carbonate aquifer in Green Creek, in northeastern Ohio 66
4-20 Distribution of chloride and oil and gas wells in Alum Creek basin, Ohio 67
4-21 Areas of ground-water pollution in Alum Creek basin, Ohio 68
4-22 Typical ground-water temperatures (°F) 70
4-23 Summer stream temperatures (°F) 70
4-24 Flow-duration curves for selected Ohio streams 71
4-25 The water table generally conforms to the surface topography 72
4-26 Aquifer A is unconfined and aquifers B and C are confined, but water may leak
through confining units to recharge adjacent water-bearing zones 74
4-27 The generalized direction of ground-water movement can be determined by
means of the water level in three wells of similar depth 75
4-28 A potentiometric surface map representing the hydraulic gradient 75
4-29 Graphical explanation of Darcy's Law 77
4-30 Using Darcy's Law to estimate underflow in an aquifer 78
4-31 Long-term ground-water hydrographs show that the water level fluctuates in
response to differences between recharge and discharge 78
4-32 Using Darcy's Law to calculate the quantity of leakage from one aquifer
to another 79
4-33 Using ground-water velocity calculations, it would require nearly six years for a
contaminant to reach the downgradient well under the stated conditions 79
4-34 Long-term ground-water hydrographs show that the water level fluctuates in
response to differences between recharge and discharge 81
4-35 Cones of depression in unconfined and confined aquifers 81
4-36 Overlapping cones of depression result in more drawdown than would be the
case for a single well 82
4-37 Values of transmissivity based on specific capacity are commonly too small
because of well construction details 83
5-1 Volume of water stored per foot of well casing for different diameter casings ... 87
5-2 Time required for recovery after slug of water removed 87
5-3 Typical multiwell installations 90
5-4 Schematic diagram of a multilevel sampling device 91
5-5 Single (a) and multiple (b) installation configurations for an air-lift sampler 91
5-6 Well developments with compressed air 93
5-7 The effects of high-velocity jetting used for well development through
openings in a continuous-slot well screen 93
5-8 Well development by back-flushing with water 93
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Figures (continued)
Number Page
5-9 Typical well protector installation 94
5-10 Cross-sectional views of (a) split spoon and (b) Shelby tube samplers 99
5-11 East-west cross section across Rock River Valley at Roscoe 100
5-12 Locations and TCE concentrations for temporary monitoring wells at
Roscoe, Illinois 101
5-13 Location of monitoring well nests and cross-section A-A'at Roscoe, Illinois . 102
5-14 Cross-section A-A' through monitoring nests 2, 3, and 4, looking in the
direction of ground water flow 104
5-15 General area of known TCE contamination 105
6-1 Steps and sources of error in ground-water sampling 109
6-2a Example of well purging requirement estimating procedure 112
6-2b Percentage of aquifer water versus time for different transmissivities 112
6-3 Schematic diagrams of common ground-water sampling devices 116
6-4 Matrix of sensitive chemical constituents and various sampling mechanisms ... 117
6-5 Generalized ground-water sampling protocol 117
6-6 Generalized flow diagram of ground-water sampling steps 119
6-7 Sample chain of custody form 124
7-1 Divergence from predicted direction of ground water 128
7-2 Example of water particle (and tracer) travel time calculation 128
7-3 Variations in ground-water flow and distribution of tracer due to
hydrodynamic dispersion 129
7-4 Movement by molecular diffusion 129
7-5 Determining the direction of ground-water flow 130
7-6 Common configurations and uses for ground-water tracing 131
7-7 Results of tracer tests at the Sand Ridge State Forest, Illinois 135
7-8 Tracer concentration at sampling well, C, measured against tracer
concentration at input, C0 136
7-9 Incomplete saturation of aquifer with tracer 136
7-10 Breakthrough curves for conservative and nonconservative tracers 137
7-11 Results of field test using a hot water tracer 137
7-12 The effect of pH on rhodamine WT 140
7-13 A comparison of the results of three simultaneous tracer tests in a karst system 140
7-14 Average annual tritium concentration of rainfall and snow for Arizona,
Colorado, New Mexico, and Utah 145
8-1 Typical ground-water contamination scenario. Several water supply
production wells are located downgradient of a contaminant source. The
geology is complex 149
8-2 Possible contaminant transport model grid design for the situation
shown in 8-1 149
8-3 Example of plots prepared with the Jacob's approximation of the Theis
analytical solution to well hydraulics in an artesian aquifer 159
8-4 Mathematical validation of a numerical method of estimating drawdown, by
comparison with an analytical solution 159
8-5 Location map for Lakewood Water District wells contaminated with volatile
organic chemicals 160
8-6 Geologic logs for Lakewood Water District wells contaminated with volatile
organic chemicals 161
8-7 Schematic illustrating the mechanism by which a downgradient source may
contaminate a production well, and by which a second well may isolate the
source through hydraulic interference 161
8-8 Location map for Chem-Dyne Superfund Site 162
8-9 Chem-Dyne geologic cross section along NNW-SSE axis 164
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Figures (continued)
Number Page
8-10 Chem-Dyne geologic cross section along WSW-ENE axis 165
8-11 Shallow well ground-water contour map for Chem-Dyne 166
8-12 Typical arrangement of clustered, vertically-separated wells installed
adjacent to Chem-Dyne and the Great Miami River 167
8-13 Estimates of transmissivity obtained from shallow and deep wells during
Chem-Dyne pump test 168
8-14 Distribution of total volatile organic chemical contamination in shallow
wells at Chem-Dyne during October, 1983 sampling 169
8-15 Distribution of tetrachloroethane in shallow wells at Chem-Dyne during
October, 1983 sampling 171
8-16 Distribution of trichloroethane in shallow wells at Chem-Dyne during
October, 1983 sampling 172
8-17 Distribution of trans-dichloroethene in shallow wells at Chem-Dyne during
October, 1983 sampling 173
8-18 Distribution of vinyl chloride in shallow wells at Chem-Dyne during
October, 1983 sampling 174
8-19 Distribution of benzene in shallow wells at Chem-Dyne during
October, 1983 sampling 175
8-20 Distribution of chloroform in shallow wells at Chem-Dyne during
October, 1983 sampling 176
8-21 General relationship between site characterization costs and clean-up costs
as a function of the characterization approach 177
8-22 Average price per category for ground-water models from the International
Ground Water Modeling Center 178
8-23 Price ranges for IBM-PC ground-water models available from various sources 179
9-1 Generalized geologic map of a glaciated area along the Souris River Valley in
central North Dakota 186
9-2 Generalized geologic cross section of the Souris River Valley based on
driller's log 186
9-3 Geologic cross section of the Souris River Valley based on detailed logs
of test holes 188
9-4 Schematic of general features of the Columbia Plateau region 188
9-5 Schematic of general features of the Piedmont and Blue Ridge region 190
9-6 Schematic of general features of the Gulf Coastal Plain 190
9-7 Schematic of general features of the Colorado Plateau and Wyoming
Basin region 192
9-8 Schematic of general features of the Nonglaciated Central region 192
9-9 Schematic of general features of the High Plains region 193
9-10 Schematic of general features of the Glaciated Central region 193
9-11 Dissolved solids concentrations in ground water used for drinking in the
United States 194
9-12 Areal extent of glacial deposits in the United States 199
9-13 Dip and strike symbols commonly shown on geologic maps 199
9-14 Cross sections of normal, reverse, and lateral faults 201
XI
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Tables
Number Page
1-1 Representative Ranges for Inorganic Constituents in Leachate from
Sanitary Landfills 3
1-2 Sources of Ground-Water Contamination 6
1-3 Range of Values of Hydraulic Conductivity 9
1-4 Exchange Capacities of Minerals and Rocks 15
3-1 Estimated Volumes of Water or Air Required to Completely Renovate
Subsurface Material that Contained Hydrocarbons at Residual Saturation 47
4-1 Selected Values of Porosity, Specific Yield, and Specific Retention 74
5-1 Advantages and Disadvantages of Selected Drilling Methods for
Monitoring Well Construction 96
6-1 Suggested Measurements for Ground-Water Monitoring Programs 109
6-2 Composition of Selected Sealing and Drilling Muds 111
6-3 Recommendations for Flexible Materials in Sampling Applications 114
6-4 Recommended Sample Handling and Preservation Procedures for a
Detective Monitoring Program 121
6-5 Field Standard and Sample Spiking Solutions 122
7-1 Comparison of Microbial Tracers 138
7-2 Measured Sorption of Dyes on Bentonite Clay 140
7-3 Sensitivity and Minimum Detectable Concentrations for the Tracer Dyes 141
7-4 Some Simple Compounds Which are Soluble in Water 143
7-5 Gases of Potential Use as Tracers 143
7-6 Properties of Fluorocarbon Compounds 144
7-7 Commonly Used Radioactive Tracers for Ground-Water Studies 145
8-1 Natural Processes that Affect Subsurface Contaminant Transport 152
8-2 Chem-Dyne Pump Test Observation Network 170
8-3 Conventional Approach to Site Characterization Efforts 177
8-4 State-of-the-Art Approach to Site Characterization Efforts 177
8-5 State-of-the-Science Approach to Site Characterization Efforts 177
8-6 Screening-Level Questions for Mathematical Modeling Efforts 181
8-7 Conceptualization Questions for Mathematical Modeling Efforts 181
8-8 Sociopolitical Questions for Mathematical Modeling Efforts 182
9-1 Geologist's Log of a Test Hole, Souris River Valley, North Dakota 187
9-2 Generalized Geologic Logs of Five Test Holes, Souris River Valley,
North Dakota 187
9-3 Geologic Time Scale 201
XII
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Preface
Background and Regulatory Objectives
Because contamination of ground water has occurred in every state and is being detected with
increasing frequency, regulatory agencies and courts have been developing guidelines, laws
and rules to protect this resource.
Ground-water quality laws deal with both the prevention of ground-water contamination and
assigning responsibility for ground-water protection or cleanup and legal liability for damages
where contamination has occurred. Provisions aimed at prevention of contamination regulate the
conduct of activities which could have the effect of polluting ground-water or posing risks to
human health.
Other statutory provisions call for government or private party response to incidents of
contamination: they also may assign penalties or other legal liability to polluters. The operation
of these provisions is generally triggered by the release of certain harmful substances, identified
by statute or government regulation, into the environment.
There is no federal law or program that directly and exclusively addresses control of ground-
water pollution. However, EPA administers a number of federal environmental laws with varying
requirements that do not exclusively address ground water, but do affect ground-water quality.
Among these are the Clean Water Act (CWA), the Resource, Conservation and Recovery Act
(RCRA), the Safe Drinking Water Act (SDWA) and the Comprehensive Environmental
Response, Compensation, and Liability Act (CERCLA), commonly referred to as Superfund.
Laws such as these and regulatory programs developed for their implementation have multiple
purposes and objectives, including protection of land and surface water quality.
Two other laws which indirectly relate to ground-water quality are the Federal Insecticide,
Fungicide and Rodenticide Act (FIFRA) and the Toxic Substances Control Act (TSCA). This
legislation regulates the production, use and disposal of specific chemicals possessing an
unacceptably high potential for contaminating ground water when released to the subsurface.
In addition, EPA has issued a policy document entitled "A Ground-Water Protection Strategy."
This strategy embraces goals to: 1) foster stronger state programs for ground-water protection
through existing Federal grant programs and provision of technical assistance; 2) study
inadequately addressed problems of ground-water contamination; and 3) strengthen the
"internal ground-water organization" within EPA by establishing an Office of Ground-Water
Protection.
Many states, also, have acted in the last several years to address the problem. Prevention of
ground-water contamination is the major thrust of most state programs. Elements of state
prevention programs include developing background data on ground-water resources,
establishing monitoring programs, and in some instances establishing permit and other
regulatory requirements to control pollution discharges into aquifers. States have also enacted
preventive legislation paralleling RCRA and SDWA, in order to qualify for federal delegation of
authority under those acts.
States have also committed funds to the cleanup of hazardous waste pollution, including
ground-water contamination. Most often, this state funding is provided as a condition of federal
Superfund financing of the cleanup of priority hazardous waste sites in the state.
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Summary of Federal Laws and Programs
CWA is one of the most far-reaching federal pollution control laws ever enacted. The Act has
application to ground-water quality control in several ways. To the extent that surface and
ground-water systems are hydrologically connected, protection of surface water quality
beneficially affects ground water. Also, funding has been provided to states for water quality
management planning and implementation, which includes ground water. In addition, where
CWA funds are used to construct municipal sewage treatment plants using land application
techniques, the municipalities are required to design the plants to ensure protection of ground
water.
RCRA was enacted in 1976 after threats to human health and the environment posed by toxic
and hazardous wastes had become matters of real public concern. The specific impetus for
RCRA's passage was Congressional concern for the special dangers caused by unsound waste
disposal practices, mainly in landfills and open generation, transportation, treatment, storage and
disposal, as well as underground storage tanks.
SOWA was passed by Congress in 1974 to respond to accumulating evidence during the
1970's that called attention to the health threat posed by unsafe levels of contaminants in public
drinking water supplies. Since about one-half of the nation's drinking water is drawn from
underground sources, SOWA has obvious application to ground-water quality, although it
applies to surface waters as well. The SDWA, as amended, provides protection to ground water
through drinking water standards, sole source aquifer designation, protection of wellheads and
the underground injection control program. FIFRA, first passed in 1947 and substantially
amended in 1972, requires that before marketing a pesticide, the manufacturer must secure a
registration of the product from EPA. In determining whether to issue a registration, EPA must
find that the pesticide will not cause "unreasonable adverse effects on the environment" if used
normally. FIFRA also imposes labeling and data reporting requirements on pesticide
manufacturers. The Act authorizes EPA to suspend or cancel the registration of a pesticide
where adverse environmental effects are shown to result from its use.
TSCA was enacted in 1976 in an effort to minimize risks to public health and the environment
posed by the introduction into commercial use of a rapidly increasing number of chemical
substances. To enable EPA to monitor the marketing of new chemicals, TSCA requires
manufacturers to submit pre-manufacture notices on new chemical substances. EPA is
authorized to take a variety of steps to protect against harmful effects caused by the
introduction or unrestricted use of new chemicals. Such steps taken by EPA under TSCA
include publication of the chemical inventory, which is a currently maintained list of all chemical
substances manufactured or processed in the U.S., as well as information gathering authority,
permitting access to manufacturing data which could assist in the development of source
inventories for ground-water protection planning or investigation.
CERCLA was passed in 1980 to respond to the notorious Love Canal incident, which focused
Congressional attention on the serious and widespread health threats posed by abandoned
hazardous waste disposal sites. Congress established Superfund to enable the federal
government to undertake prompt cleanup of especially dangerous abandoned sites, and later to
seek reimbursement from the responsible parties. CERCLA applies cleanup, funding and liability
provisions as triggered by a release or threat of release of a hazardous substance from a
facility.
EPA's Ground-Water Protection Strategy sets out a "policy framework" to guide its programs
affecting ground water. This framework involves classification of ground waters. Class I are
those "special ground waters" in need of special protection because they are irreplaceable
sources of drinking water or are otherwise ecologically vital, and are highly vulnerable to
contamination because of hydrogeologic factors. With RCRA authority, EPA will ban siting of
disposal facilities above these ground waters. The Agency will also continue to use the
immediacy of a threat to ground water as a factor in selecting sites for Superfund cleanup.
Further, EPA is considering developing special permit conditions for the underground injection
control program to protect these waters. Class II ground waters are those used or potentially
available for drinking water, though less vulnerable than Class I aquifers. Class II aquifers
presently account for most of the country's ground water. As to these waters, EPA may impose
XIV
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facility siting restrictions under RCRA. Class III ground waters are classified as waters that are
not potential sources of drinking water and are of limited beneficial use.
Purpose
The subsurface environment of ground water is characterized by a complex interplay of
physical, geochemical and biological forces that govern the release, transport and fate of a
variety of chemical substances. There are literally as many varied hydrogeologic settings as
there are types and numbers of contaminant sources. In situations where ground-water
investigations are most necessary, there are frequently many variables of land and ground-
water use and contaminant source characteristics which cannot be fully characterized.
The impact of natural ground-water recharge and discharge processes on distributions of
chemical constituents is understood for only a few types of chemical species. Also, these
processes may be modified by both natural phenomena and man's activities so as to further
complicate apparent spatial or temporal trends in water quality. Since so many climatic,
demographic and hydrogeologic factors may vary from place to place, or even small areas
within specific sites, there can be no single "standard" approach for assessing and protecting
the quality of ground water that will be applicable in all cases.
Despite these uncertainties, investigations are under way and they are used as a basis for
making decisions about the need for, and usefulness of, alternative corrective and preventive
actions. Decision makers, therefore, need some assurance that elements of uncertainty are
minimized and that hydrogeologic investigations provide reliable results.
A purpose of this document is to discuss measures that can be taken to ensure that
uncertainties do not undermine our ability to make reliable predictions about the response of
contamination to various corrective or preventive measures.
EPA conducts considerable research in ground water to support its regulatory needs. In recent
years, scientific knowledge about ground-water systems has been increasing rapidly.
Researchers in the Office of Research and Development have made improvements in
technology for assessing the subsurface, in adapting techniques from other disciplines to
successfully identify specific contaminants in ground water, in assessing the behavior of certain
chemicals in some geologic materials and in advancing the state-of-the-art of remedial
technologies.
An important part of EPA's ground-water research program is to transmit research information
to decision makers, field managers and the scientific community. This publication has been
developed to assist that effort and, additionally, to help satisfy an immediate Agency need to
promote the transfer of technology that is applicable to ground-water contamination control and
prevention.
The need exists for a resource document that brings together available technical information in
a form convenient for ground-water personnel within EPA and state and local governments on
whom EPA ultimately depends for proper ground-water management. The information
contained in this handbook is intended to meet that need. It is applicable to many programs that
deal with the ground-water resource. However, it is not intended as a guidance or support
document for a specific regulatory program.
GUIDANCE DOCUMENTS ARE AVAILABLE FROM EPA AND MUST BE CONSULTED TO
ADDRESS SPECIFIC REGULATORY ISSUES.
xv
-------
Acknowledgments
Many individuals contributed to the preparation and review of this handbook. The document was
prepared by JACA Corporation for EPA's Robert S. Kerr Environmental Research Laboratory,
Ada, OK, and the Center for Environmental Research Information, Cincinnati, OH. Contract
administration was provided by the Center for Environmental Research Information, Cincinnati,
OH.
Authors:
Michael Barcelona - Illinois State Water Survey, Champaign, IL
Joseph F. Keely - EPA-RSKERL, Ada, OK
Wayne A. Pettyjohn - Oklahoma State University, Stillwater, OK
Allen Wehrmann - Illinois State Water Survey, Champaign, IL
Reviewers and Other Contributors:
Edwin F. Barth - EPA-OERR, Washington, DC
Stuart 2. Cohen - EPA-OTS, Washington, DC
Stephen Cordle - EPA-OEPER, Washington, DC
Mary Doyle - University of Arizona, Tucson, AZ
Gerald Grisak - Intera Technologies, Austin, TX
Kenneth Jennings - EPA-OWPE, Washington, DC
Jerry N. Jones - EPA-RSKERL, Ada, OK
Lowell E. Leach - EPA-RSKERL, Ada, OK
Joan Middleton - EPA-OSWER, Washington, DC
Marion R. Scalf - EPA-RSKERL, Ada, OK
Jerry T. Thornhill - EPA-RSKERL, Ada, OK
Calvin H. Ward - Rice University, Houston, TX
Contract Project Officer:
Carol Grove - EPA-CERI, Cincinnati, OH
XVI
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CHAPTER 1
GROUND-WATER CONTAMINATION
1.1 Definitions
Contaminant is defined by the Safe Drinking Water
Act as "any physical, chemical, biological, or
radiological substance or matter in water." Freeze
and Cherry (1979) define as contaminants "all solutes
introduced into the hydrologic environment as a result
of man's activities regardless of whether or not the
concentrations reach levels that cause significant
degradation of water quality." For them, "pollution is
reserved for situations where contaminant
concentrations attain levels that are considered to be
objectionable." Miller (1980) used a very similar
definition: "Ground-water contamination is the
degradation of the natural quality of ground water as a
result of man's activities." According to Matthess
(1982), "boundaries of polluted ground-water zones
can be defined as the lines at which the concentration
of all pollutants have fallen below the maximum
permissible concentration for potable water, or where
all water properties have taken on the normal values
of the environment concerned."
Much current research is being devoted to defining
just what "normal" ground-water quality is, or how it
can best be defined. Ground water which naturally
contains objectionable amounts of dissolved
substances can properly be considered contaminated,
as well as polluted; however, most regulatory
functions focus on human activities which artificially
introduce contaminants into the ground.
1.2 The Extent of Ground-Water
Contamination
Contrary to what many people believe, ground-water
contamination is not a new problem. Early
investigations of ground-water contamination are
abundant in scientific literature. The classic work of
Dr. John Snow in 1854 (Prescott and Horwood, 1935;
Mailman and Mack, 1961) first linked the
contamination of wells by cholera to seepage from
earth privy vaults even before the discovery of the
microorganisms responsible for the disease. By 1959,
a European publication (Michels et a/.) cited 60 cases
in which ground waters had become contaminated
with petroleum products.
LeGrand, in his 1965 paper entitled "Patterns of
Contaminated Zones of Water in the Ground"
recognized the difficulty in predicting the spatial
extent of a contaminated zone because of a number
of interrelated factors including:
"...the great variety of waste materials, their
range in toxicity and adverse effects; man's
variable pattern of waste disposal and of
accidental release of contaminants in the
ground; man's variable pattern of water
development from wells; behavior of each
contaminant in the soil, water, and rock
environment; ranges in geologic and
hydrologic conditions in space; and ranges in
hydrologic conditions in time."
Additional problems include the fact that many
potentially hazardous contaminants are colorless,
odorless, and tasteless, and therefore difficult to
detect by passive means. Many of the synthetic
organic chemicals require sophisticated, expensive,
sampling and analytical techniques burdening
detection efforts.
It has been estimated that it will take 4 to 5 years to
complete just one round of organic compound testing
of the 3,400 public water supply wells in Illinois alone,
given the present availability of personnel and
laboratory facilities (Illinois EPA, 1986). Such an effort
does not include the estimated 500,000 private wells
in the State.
An assessment of the extent and severity of
contamination is further complicated by the almost
exponential growth of the synthetic organic chemistry
industry in the U.S. since the early '40s (Figure 1-1).
At least 63,000 synthetic organic chemicals are in
common industrial and commercial use in the U.S.
and this number continues to grow by approximately
500 to 1,000 new compounds every year (Epstein,
1979; U.S. EPA, 1979). Also, the human health
effects of many of these chemicals, particularly over
long periods of time at low exposure levels, is not
known. It will take years to conduct the research
necessary to properly test all these compounds and
then be able to factor the results into a complete
contamination assessment.
-------
Figure 1-1
Growth of the synthetic organic chemical
industry in the United States (from Senkan
and Stauffer, 1981).
10"
10"
10"
10s
I
1915 1925 1935
1945 1955
Year
1965
1975
Though it has now been estimated that approximately
1 percent of the economically producible ground
waters in the United States are contaminated (Lehr,
1982; Gass, 1980; Office of Technology Assessment,
1984), this estimate may not convey the problems
associated with the coincidence of contamination and
ground-water use. While on the whole, much of the
ground water in the U.S. has not been affected by
contamination, areas known to be contaminated are
often densely populated areas where ground water is
heavily used and depended upon as a drinking water
source.
The presence of over 200 chemical substances in
ground water has been documented (OTA, 1984).
This number includes approximately 175 organic
chemicals, over 50 inorganic chemicals (metals,
nonmetals, and inorganic acids), and radionuclides.
Many of these chemicals occur naturally in ground
water, especially minerals dissolved from geologic
earth materials in contact with the water. Many others
have been introduced to the ground-water system
by humans.
The detection of these substances has been biased
by sampling and analytical limitations as well as the
nature of the specific investigations which prompted
the sampling and analysis to be conducted. The two
most common circumstances under which
substances (including naturally occurring minerals)
have been detected in ground water are (a) regulatory
compliance (e.g., Safe Drinking Water Act monitoring
of public water supplies and Resource Conservation
and Recovery Act (RCRA) monitoring at hazardous
waste facilities); and (b) response to perceived quality
problems, primarily citizen complaints. However,
regulatory agencies have not historically sampled and
analyzed for a wide range of potential contaminants,
particularly synthetic organic chemicals, unless
specific problems are suspected.
A study of the ground-water quality data base
maintained in Illinois (O'Hearn and Schock, 1984)
found that compliance monitoring for drinking water
standards forms the basis for much of the 21,000
samples and 423,000 analytical determinations in this
data base. However, less than one-tenth of 1
percent of all the samples in the data base had been
analyzed for even a general indicator of organic
contamination, total organic carbon (TOC).
Efforts have been made in recent years to assess the
occurrence of organic chemicals in ground-water
supplies. A survey conducted by the U.S. EPA, the
Ground Water Supply Survey (GWSS), provided
information on the frequency with which VOCs were
detected in 466 randomly selected public ground-
water supply systems (Westrick ef a/., 1983). One or
more volatile organic chemicals (VOCs) were
detected in 16.8 percent of small systems and 28.0
percent of large systems sampled. The two VOCs
found most often in this survey were trichloroethylene
(TCE) and tetrachloroethylene (PCE). Two or more
VOCs were found in 6.8 percent and 13.4 percent of
the samples from small and large systems,
respectively.
1.3 General Mechanisms of Ground-
Water Contamination
Contaminant releases to ground water can occur by
design, by accident, or by neglect. Most ground-
water contamination incidents involve substances
released at or only slightly below the land surface.
Consequently, it is shallow ground water which is
affected initially by contaminant releases. In general,
shallow ground-water resources are considered
more susceptible to surface sources of contamination
than deeper ground-water resources. There are at
least four ways by which ground-water
contamination occurs: infiltration, direct migration,
interaquifer exchange, and recharge from surface
water. A general discussion of each of these
mechanisms follows.
1.3.1 Infiltration
Contamination by infiltration is probably the most
common ground-water contamination mechanism. A
portion of the water which has fallen to the earth
slowly infiltrates the soil through pore spaces in the
-------
soil matrix. As the water moves downward under the
influence of gravity, it dissolves materials with which it
comes into contact. Water percolating downward
through a contaminated zone can dissolve
contaminants, forming leachate. Depending on the
composition of the contaminated zone, the leachate
formed can contain a number of inorganic and
organic constituents. Table 1-1 gives a general
indication of the composition of leachate that has
been found beneath sanitary landfills. The leachate
will continue to migrate downward under gravity's
influence until the saturated zone is reached. Once
the saturated zone is contacted, horizontal and
vertical spreading of the contaminants in the leachate
will occur in the direction of ground-water flow
(Figure 1-2). This process can occur beneath any
surface or near-surface contaminant source
exposed to the weather and the effects of infiltrating
water.
Table 1-1 Representative
Constituents in
Landfills.
Parameter
K*
Na*
Ca2*
Mg*
ci-
scv-
Alkalinity
Fe (total)
Mn
Cu
Ni
Zn
Pb
Hg
N03-
NH4
P as P04
Organic nitrogen
Total dissolved organic carbon
COD (chemical oxidation demand)
Total dissolved solids
pH
Ranges for Inorganic
Leachate from Sanitary
Representative Range
(mg/l)
200-1,000
200-1,200
100-3,000
100-1,500
300-3,000
10-1,000
500-10,000
1-1,000
0.01-100
<10
0.01-1
0.1-100
<5
<0.2
0.1-10
10-1,000
1-100
10-1,000
200-30,000
1,000-90,000
5,000-40,000
4-8
Source: Freeze and Cherry, 1979.
1.3.2 Direct Migration
Contaminants can migrate directly into ground water
from below-ground sources (e.g., storage tanks,
pipelines) which lie within the saturated zone.
Leachate formation and downward movement through
the unsaturated zone need not occur prior to
contamination of nearby ground water. Much greater
concentrations of contaminant may occur because of
the continually saturated conditions. Storage sites and
landfills excavated to a depth near the water table
also may permit direct contact of contaminants with
ground water. Another direct entry of contaminants
from the surface to the ground-water system may
be from the vertical leakage of contaminants through
the seals around well casings or through improperly
abandoned wells, or as a result of contaminant
disposal through deteriorated or improperly
constructed wells.
7.3.3 Interaquifer Exchange
Contaminated ground water can mix with
uncontaminated ground water through a process
known as interaquifer exchange in which one water-
bearing unit "communicates" hydraulically with
another. This is most common in bedrock aquifers
where a well penetrates more than one water-
bearing formation to provide increased yield. Each
water-bearing unit will have its own head potential,
some greater than others. When the well is not being
pumped, water will move from the formation with the
greatest potential to formations of lesser potential. If
the formation with the greater potential contains
contaminated or poorer quality water, the quality of
water in another formation can be degraded.
Similar to the process of direct migration, old and
improperly abandoned wells with deteriorated casings
or seals are a potential contributor to interaquifer
exchange. Vertical movement may be induced by
pumping or may occur under natural gradients. For
example, in Figure 1-3, an improperly abandoned
well formerly tapping only a lower uncontaminated
aquifer suffers from a corroded casing. This allows
water from an overlying contaminated zone to
communicate directly with the lower aquifer. The
pumping of a nearby well tapping the lower aquifer
creates a downward gradient between the two
water-bearing zones. As pumping continues,
contaminated water migrates through the lower
aquifer to the pumping well. Downward migration of
the contaminant may also occur through the aquitard
(confining layer) separating the upper and lower
aquifers. However, the rate of movement through the
aquitard is often much slower than the rate at which
contaminants move through the direct connection of
an abandoned well.
-------
Figure 1-2 Plume of leachate migrating from a sanitary landfill on a sandy aquifer using contours of chloride
concentration (from Freeze and Cherry, 1979).
Horizontal Scale
Chloride concentration, mg/l
Standpipetip
Piezometer tip
Multi-level sampling point
Water table
Clay
Flow direction
Figure 1-3 Vertical movement of contaminants along an
old, abandoned, or improperly constructed
well (from Deutsch, 1961).
Abandoned Well
Disposal Pond (Corroded Casing)
Upper Aquifer
Contaminated Water
To Municipal
Supply
1.3.4 Recharge from Surface Water
Normally, ground water moves toward or
"discharges" to surface water bodies (see discussion,
Chapter 4). Occasionally, however, the hydraulic
gradient is such that surface water has a higher
potential than ground water (such as during flood
stages), causing a reversal in flow. Contaminants in
the surface water can then enter the ground-water
system.
Reversal of flow can also be caused by pumping
(Figure 1-4). Lowering the ground-water level to a
level near a surface water body can induce leakage
through the stream or lake bed. Contamination of a
glacial sand and gravel aquifer by organic compounds
present in an adjacent river in such a manner has
been documented (Schwarzenbach et al., 1983).
1.4 Sources of Ground-Water
Contamination
A wide variety of ground-water contamination
sources have been identified. As previously
mentioned, contaminant releases to ground water can
occur by design, by accident, or by neglect. The
Office of Technology Assessment (OTA, 1984)
grouped 33 types of ground-water contamination
sources into six major categories (Table 1-2) based
on the general nature of the contaminating activity. A
number of these sources are depicted in Figure 1-5.
Category 1 includes sources that are intentionally
designed to discharge substances. Subsurface
percolation systems, such as septic tanks and cess
pools, injection wells, and land application of
wastewater or sludges fall within this category. Such
-------
Figure 1-4 Contaminated water induced to flow from
surface water to ground water by pumping
(from Miller, 1980).
Contaminated
Surface Water
systems are primarily designed to use the natural
capacity of the soil materials to degrade wastewaters.
Injected wastewaters are often placed in unusable
zones to be assimilated with poor quality ground
water of natural origin. Septic tanks and cess pools
have been estimated to discharge the largest volume
of wastewater into the ground and are the most
frequently reported source of ground-water
contamination (Miller, 1980).
Injection wells are another major potential source of
contamination. Although injection wells can be
constructed and operated properly, contamination of
ground water can occur in several ways (EPA, 1979):
o Faulty well construction (e.g., drilling and casing)
o The forcing upward of pressurized fluids into
nearby wells and aquifers, and faults and
fractures of confining beds
o The migration of fluids into hydrologically
connected usable aquifers
o Faulty well closing.
The depths and operating procedures used in waste
injection generally make monitoring and leak
prevention very difficult to validate.
Land application is a popular, inexpensive alternative
for wastewater and sludge treatment. The U.S. EPA
(1983) estimated that 40 to 50 percent of the
municipal sludge generated every year is applied to
the land.
Category II includes sources that are designed to
store, treat, or dispose of substances but are not
designed to release contaminants to the subsurface.
Landfills, open dumps, local residential disposal,
surface impoundments, waste tailings and piles,
materials stockpiles, graveyards, aboveground and
underground storage tanks, containers, open burning
sites, and radioactive disposal sites all fall into this
broad category. It is important to note here that while
a number of sources in this category are considered
"waste" sources (e.g., landfills, dumps,
impoundments, etc.), many others are "non-waste"
related sources. Storage tanks, stockpiles, and a
variety of containers with residues of commercial
products have been found to contribute contaminants
to ground water.
Category III consists of sources designed to retain
substances during transport or transmission. Such
sources primarily consist of pipelines and material
transport or transfer operations. Contaminant releases
generally occur by accident or neglect; for example,
as a result of pipeline breakage or a traffic accident.
Again, most substances which would be subject to
release from sources within this category are not
wastes but raw materials or products to be used for
some beneficial purpose.
Category IV includes those sources discharging
substances as a consequence of other planned
activities. This category contains a number of
agriculturally related sources such as irrigation return
flows, feedlot operations, and pesticide and fertilizer
applications. A number of sources related to urban
activities such as highway desalting, urban runoff, and
atmospheric deposition are included. Surface and
underground mine-related drainage also fall within
this category.
Category V comprises sources providing conduits or
inducing discharge through altered flow patterns. For
the most part, such sources are unintentional
ground-water contamination sources and include
water, oil, and gas production wells, monitoring wells,
exploration holes, and construction excavations. The
potential to contaminate ground water from production
wells stems from poor installation and operation
methods, and incorrect plugging or abandonment
procedures. Such practices create opportunities for
cross-contamination by vertical migration of
contaminants.
Finally, Category VI includes naturally occurring
sources whose discharge is created or made worse
by human activity. Ground-water/surface water
interactions, described in the previous section, and
salt-water intrusion or upconing (ground-water
movement upward as a result of pumpage) provide
the basis for this category. Withdrawals significantly in
excess of recharge can affect ground-water quality.
Salt water intrusion in coastal areas and brine-water
upconing from deeper formations in inland areas can
occur when pumpage exceeds the aquifer's natural
recharge rate.
Contaminant releases are also referred to as
originating from point or nonpoint sources. Point
sources are those which release contaminants from a
discrete geographic location. Examples include
leaking underground storage tanks, septic systems,
and injection wells. Nonpoint contamination situations
-------
Table 1-2 Sources of Ground-Water Contamination (from OTA, 1984).
Category I—Sources designed to discharge substances
Subsurface percolation (e.g., septic tanks and cesspools)
Injection Wells
Hazardous waste
Non-hazardous waste (e.g., brine disposal and drainage)
Non-waste (e.g., enhanced recovery, artificial recharge, solution
mining, and in-situ mining)
Land application
Wastewater (e.g., spray irrigation)
Wastewater byproducts (e.g., sludge)
Hazardous waste
Non-hazardous waste
Category II—Sources designed to store, treat, and/or dispose of
substances; discharge through unplanned release
Landfills
Industrial hazardous waste
Industrial non-hazardous waste
Municipal sanitary
Open dumps, including illegal dumping (waste)
Residential (or local) disposal (waste)
Surface impoundments
Hazardous waste
Non-hazardous waste
Waste tailings
Waste piles
Hazardous waste
Non-hazardous waste
Materials stockpiles (non-waste)
Graveyards
Animal burial
Aboveground storage tanks
Hazardous waste
Non-hazardous waste
Non-waste
Underground storage tanks
Hazardous waste
Non-hazardous waste
Non-waste
Containers
Hazardous waste
Non-hazardous waste
Non-waste
Open burning and detonation sites
Radioactive disposal sites
Category III—Sources designed to retain substances during
transport or transmission
Pipelines
Hazardous waste
Non-hazardous waste
No n-waste
Materials transport and transfer operations
Hazardous waste
Non-hazardous waste
Non-waste
Category IV—Sources discharging substances as consequence
of other planned activities
Irrigation practices (e.g., return flow)
Pesticide applications
Fertilizer applications
Animal feeding operations
De-icing salts applications
Urban ruhnoff
Percolation of atmospheric pollutants
Mining and mine drainage
Surface mine-related
Underground mine-related
Category V—Sources providing conduit or inducing discharge
through altered flow patterns
Production wells
Oil (and gas) wells
Geothermal and heat recovery wells
Water supply wells
Other wells (non-waste)
Monitoring wells
Exploration wells
Construction excavation
Category VI—Naturally occurring sources whose discharge is
created and/or exacerbated by human activity
Groundwater—surface water interactions
Natural leaching
Salt-water intrusion/brackish water upconing (or intrusion and
other poor-quality natural water)
-------
Figure 1-5 Sources of ground-water contamination (from Geraghty and Milter, 1985).
Intentional
Input
Unintentional
Input
Ground-Water
Movement
are more extensive in area and diffuse in nature. It is
therefore difficult to trace contaminants from nonpoint
sources back to their origin. Agricultural activities (i.e.,
application of pesticides and fertilizers), urban runoff,
and atmospheric deposition are potential nonpoint
contaminant sources.
1.5 Movement
Ground Water
of Contaminants in
7.5.7 Contaminant Migration
In broad terms, three processes govern the migration
of chemical constituents in ground water: (1)
advection, movement caused by the flow of ground
water; (2) dispersion, movement caused by the
irregular mixing of waters during advection; and (3)
retardation, principally chemical mechanisms which
occur during advection.
1.5.1.1 Advection
Ground water in its natural state is constantly in
motion (advection), although in most cases it is
moving very slowly (Todd, 1980). Ground-water
movement is governed by the hydraulic principles
discussed in Chapter 4.
For example, Darcy's Law states that the flow rate
through any porous medium is proportional to the
head loss and inversely proportional to the length of
the flow path:
Q = -KxAxbi/L (1-1)
where:
Q = ground-water flow rate, in gal/d
A = cross-sectional area of flow, in ft2
hj =head loss, in feet, measured between two
points L ft apart
K = hydraulic conductivity, a measure of the
ability of the porous medium to transmit
water, in gal/d/ft2
Equation 1-1 can be rearranged in the following
manner to produce the "bulk," or what is called
Darcian, velocity:
-------
v = 7.48 (Q/A)
v = 7.48 (-K)(hi/L)
v = 7.48 (-K)(dh/dl)
(1-2)
where:
v = Darcian velocity of ground-water flow, in
ft/d
dh = the change in hydraulic head (head loss),
in ft
dl = the distance or change in position (length)
over which the head loss is measured, in ft.
The Darcian velocity assumes that flow occurs across
the entire cross section of the porous material without
regard to solid or pore spaces.
Actually, flow is limited to the pore space only, so the
actual "interstitial" flow velocity is:
Va = v/n = 7.48 K/n x dh/dl
(1-3)
where:
Va =
n =
the actual ground-water flow velocity, in
ft/d
the effective porosity, or the percent of the
porous media which consists of
interconnected pore spaces, the spaces
which contribute to ground-water flow,
unitless.
The hydraulic conductivity of a geologic formation
depends on a variety of physical factors, including
porosity, particle size and distribution, the shape of
the particles, particle arrangement (packing), and
secondary features such as fracturing and dissolution.
In general, for unconsolidated porous materials,
hydraulic conductivity values vary with particle size.
Fine-grained, clayey materials exhibit lower values of
hydraulic conductivity while coarse-grained sandy
materials normally exhibit higher conductivities. Table
1-3 shows the range of values commonly exhibited
by geologic materials.
The effective porosity is essentially an estimated
parameter because the actual measurement of the
volume of interconnected pore spaces in most porous
media has not been conducted. Effective porosity is
usually estimated as being somewhat less than the
total porosity. Total porosity is calculated from ratios
of the volumes of saturated and dry porous material.
In coarse-grained materials which drain freely, the
effective porosity is essentially equal to total porosity
and is generally defined as the ratio of the volume of
water which drains by gravity to the total volume of
saturated porous material.
Equation 1-3 (interstitial velocity) has been used for
determining the advective component of ground-
water flow and as a conservative estimate of the rate
of migration of dissolved constituents. The rate of
movement of the front of a dissolved constituent
"plume" by the process of advection can be
calculated in a similar fashion.
Figure 1-6 shows the relative concentration of a
dissolved constituent emanating from a constant
source of contamination versus distance along the
flow path. Figure 1-7 shows a similar plot for a
discontinuous contaminant source which produced a
single slug of dissolved contaminant. In both cases,
advective movement causes the dissolved constituent
to move with the ground water at the average rate
described in Equation 1-3. Considering advective
flow only, no diminution of concentration appears as a
straight line moving at the rate of ground-water flow.
Figure 1-8 shows the effect of advection on the
movement of a contaminant in a regional ground-
water flow field. Contaminants moving out of a leaking
lagoon move horizontally and vertically following the
pattern of flow established by ground water as it
moves from an upgradient area of recharge to the
zone of discharge at the river. Mechanisms
influencing the spread of contaminant in the flow field
are discussed in the following sections.
1.5.1.2 Dispersion
In natural porous materials, the pores possess
different sizes, shapes, and orientations. Similar to
stream flow, a velocity distribution exists within the
pore spaces such that the rate of movement is
greater in the center of the pore than at the edges.
Therefore, in saturated flow through these materials,
velocities vary widely across any single pore and
between pores. As a result, a miscible fluid will
spread gradually to occupy an ever increasing portion
of the flow field when it is introduced into a flow
system. This mixing phenomenon is known as
dispersion. In this sense, dispersion is a mechanism
for dilution.
Dispersion can occur both in the direction of flow and
transverse (perpendicular) to it. Dispersion caused by
microscopic changes in flow direction due to pore
space orientation is depicted in Figure 1-9a.
Macroscopic features, such as fingered lenses of
higher conductivity, are shown in Figures 1-9b and
1-9c. Solution channeling and fracturing are other
macroscopic features which may contribute to
contaminant dispersion (Figure 1-10). Careful
placement of wells is required when monitoring in
complicated geologic systems such as those shown
in Figures 1-9(b and c) and 1-10.
The effect of dispersion as a plot of relative
constituent concentration versus distance along a
flow path is shown in Figure 1-11. Notice that the
front of the dissolved constituent distribution is no
longer straight but rather appears "smeared." Some
dissolved constituent actually moves ahead of what
would have been predicted if only advection were
considered.
-------
Table 1-3 Range of Values of Hydraulic Conductivity (adapted rom Freeze and Cherry, 1979).
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-------
Figure 1-6 Movement of a concentration front by
advection only.
o
0
0
Dissolved Constituent
Average Flow
Distance >.
Figure 1-7 Movement of a dissolved constituent slug by
advection only.
o
o
0
Average Flow
Distance •
In a similar manner, the concentration of a slug of
material introduced to a flow field will appear as
shown in Figure 1-12 (a and b). The peak
concentration is reduced over time and distance. In
such a situation, the total mass of dissolved
constituent remains the same; however, a larger
volume is occupied, effectively reducing the
concentration found at any distance along the flow
path. In plane view, continuous and intermittent
sources affected by dispersion will appear as shown
in Figure 1-13.
1.5.1.3 Retardation
In ground-water contaminant transport, there are a
number of chemical and physical mechanisms which
retard, that is, delay or slow the movement of
constituents in ground water. Four general
mechanisms can retard the movement of chemical
constituents in ground water: dilution, filtration,
chemical reaction, and transformation.
Figure 1-14 illustrates the movement of a
concentration front by advection only, and with
dispersion, sorption, and biotransformation. The
combined effects of advection, dispersion, sorption,
and biotransformation on a slug of contaminant
introduced into a flow system is also shown in Figure
1-14.
Dilution does not retard the movement of ground-
water constituents. However, dilution may lessen the
severity of contamination by reducing peak
concentrations encountered in the ground-water
system. For this reason, dilution, particularly by
Figure 1-8 Effect of leakage from a lagoon on a regional flow pattern (from Geraghty and Miller, 1985).
Monitoring Well
Leaky
Lagoon
Water Table
River
Head Contours -
10
-------
Figure 1-9 Comparison of advance of contaminant influenced by hydrodynamic dispersion (adapted from Freeze and
Cherry, 1979).
Particle
Porous
Medium
Transverse
Dispersion
Longitudinal
Dispersion
Mean Flow
Mean Flow
A. Microscopic scale of a granular medium
Plug Flow
Tracer Input
Tracer
Injection Points
B. Macroscopic illustration of fingering caused by layered
beds and lenses
Coarse Lens
I Higher
> K
I Lenses
C. Spreading caused by irregular lenses.
11
-------
Figure 1-10 Flow of contaminated ground water in aquifer
with solution porosity (from Geraghty and
Miller, 1985).
Figure 1-11 Movement of a concentration front by
advection and dispersion.
Flow
o
c
g
1
o
O
15
cn
Dissolved Constituent
Front from Advection Only
Average Flow
Distance •
Figure 1-12 Movement of a dissolved constituent slug by
advection and dispersion as it moves from
time period (a) to (b).
Time Period A
u
o
Time Period B
o>
; Advection
I • Component
1 Only
Average Flow
Distance
Figure 1-13 Continuous and intermittent sources affected
by dispersion.
A. The development of a contamination plume from a
continuous point source.
Flow direction of leachate
Source
Leachate enriched ground water
Flow
B. The travel of a contaminant slug(s) from a one-time
point source or an intermittent source.
o CD
12
-------
Figure 1-14 The influence of natural processes on levels
of contaminants downgradient from
continuous and slug-release sources.
A Advection
D Dispersion
S Sorption
B Biotransformation
Distance from Continuous Contaminant Source
A + D + S
A+D+S+B
A + D
Distance from Slug-Release Contaminant Source
dispersive mechanisms, is included by many
scientists in discussions of retardation.
Filtration occurs as dissolved and solid matter are
trapped in the pore spaces of the soil and aquifer
media, clogging the pore spaces and limiting flow. As
the clogging process continues, a decrease in the
hydraulic conductivity of the material is manifested.
Chemical reactions can also take place which cause
a dissolved molecule, for example, to combine with
another such that the size of the new molecule is too
large for the pore space and mechanical filtration
occurs. Flocculation of colloidal material or gas
bubble formation may cause eventual clogging of
pore spaces resulting in a filtering effect. Microbial
activity, especially when paniculate matter and
dissolved organic materials are present together, can
enhance biological growth such that pore spaces
become blocked, hindering the movement of
dissolved constituents.
Ion exchange processes exert an important influence
on retarding the movement of chemical constituents
in ground water. In ground water systems, ion
exchange occurs when ions (electrically charged
particles) in solution displace ions associated with
geologic materials. In Figure 1-15 the originally
dissolved calcium (Ca2 + ) ion becomes bound to the
geologic media by displacing two sodium (Na + ) ions
which have less affinity for the exchange sites on the
geologic "matrix." This ion exchange process
removes constituents from the ground water and
releases others to the flow system. The calcium ion
will stay bound to the site until another ion with a
greater affinity for that site comes along or a shift in
environmental conditions (such as a change in pH)
causes the ion to release from its site. Ion exchange
capacity is very dependent on pH; metal ions, in
particular, may exchange onto geologic materials
quite readily at neutral pH (~7) but will be displaced
readily by hydrogen ions when the pH is lowered.
One major consideration in ion exchange is that the
exchange capacity of a geologic material is limited. A
measure of this capacity is quantified in a term called
"ion exchange capacity" and is defined as the
amount of exchangeable ions, in milliequivalents per
100 grams solids at pH 7. The exchange capacities of
several different subsurface materials are given in
Table 1-4 (from Matthess, 1982). Typically, clay
minerals (e.g., montmorillonite) exhibit greater cation
exchange capacities than other minerals such as
quartz (the primary component of sand). This is
because the available surface area of the clays is
often much greater than other minerals.
It is important to recognize that the exchange
capacity of a geologic material may retard
contaminant movement from a waste source for years
or even decades. However, if the source continues to
supply a highly ionized leachate, it is possible to
exceed the exchange capacity of the geologic
material, eventually allowing unretarded transport.
Changes in environmental conditions or ground-
water solution composition can also cause the release
of constituents formerly bound to the geologic
materials.
Anionic exchange in aquifer systems is not as well
understood as cationic exchange. Anions such as
sulfate, chloride, and nitrate would not be expected to
be retarded significantly by anion exchange because
most mineral surfaces in natural water systems are
negatively charged. Chloride ions may be regarded as
conservative or non-interacting ions which move
largely unretarded with the advective velocity of the
ground water. An example of the copper metal ion
(Cu2 + ) being retarded along the flow path while the
chloride ion (CI") moves unretarded is shown in
Figure 1-16.
The release of ions by exchange processes may
aggravate a pollution problem. Increases in water
hardness as a result of the displacement of calcium
and magnesium ions from geologic materials by
sodium or potassium in landfill leachate has been
documented (Hughes ef a/., 1971). The release of
aluminum to solution, in addition to calcium and
magnesium, from soils reacted with an industrial
13
-------
Figure 1-15 Ion exchange (from Geraghty and Miller,
1985).
Figure 1-16 Metal-ion movement slowed by ion exchange
(from Geraghty and Miller, 1985).
Cu
waste has also been documented (Rovers et a/.,
1976).
A number of other chemical reactions can influence
the movement of contaminants in ground water.
These include precipitation and complexation.
Chemical precipitation in waste leachates is controlled
primarily by pH and ionic concentration products.
Precipitation of metals as hydroxides, sulfides, and
carbonates is very common. Complexation involves
the formation of soluble charged or electrically neutral
complexes called ligands, which form between metal
ions and either organic or inorganic species. For
example, the complexation of Cobalt-60 ions by both
neutral and synthetic organic compounds enhanced
subsurface mobility of the radionuclide (Killey et a/.,
1984). Other metal species and organic pesticides
have been observed to travel significant distances in
ground water after the formation of soluble organic
complexes with humic substances or organic solvents
(Broadbent and Ott, 1957; Griffin and Chou, 1980;
Duguid, 1975).
Other processes which may affect contaminant
distribution and transport include volatilization as well
as a number of transformation mechanisms. The
process which occurs when a substance changes
from the liquid phase to the gaseous phase is called
volatilization. A number of organic compounds
(benzene, TCE, and many other low molecular weight
compounds) partition into and diffuse through soil gas
as a result of their low aqueous solubility and high
vapor pressure (low boiling point). Volatilization is
enhanced by low soil moisture and high air porosity
which generally occurs in coarse-textured materials
such as sand and gravel. Remote detection
techniques capable of locating subsurface volatile
organic chemical plumes by analyzing the overlying
soil gases have been devised to take advantage of
the result of volatilization (Marrin, 1985). Hydrolysis or
chemical reactions may also transform or partially
degrade some components of a waste or contaminant
mixture.
In addition, the transformation of carbonaceous and
inorganic chemicals by microorganisms with evolution
of CO2, CH4, H2, H2S, N2, NHa, and NO gases
readily occurs in many landfill and other subsurface
environments. Microbial processes may be a major
factor in the transformation of organic materials
present in ground water.
Under the appropriate circumstances pollutants can
be completely degraded to harmless products. Under
other circumstances, however, they can be
transformed to new substances that are more mobile
or more toxic than the original contaminant.
Quantitative predictions of the fate of biologically
reactive substances are at present very primitive,
particularly compared to other processes that affect
pollutant transport and fate. This situation resulted
from the ground-water community's choice of an
inappropriate conceptualization of the active
processes: subsurface biotransformations were
presumed to be similar to biotransformations known
to occur in surface water bodies. Only very recently
has detailed field work revealed the inadequacy of the
prevailing view.
As little as 5 years ago ground-water scientists
considered aquifers and soils below the zone of plant
roots to be essentially devoid of organisms capable of
transforming contaminants. However, recent studies
have shown that water-table aquifers harbor
appreciable numbers of metabolically active
microorganisms, and that these microorganisms
frequently can degrade organic contaminants. Thus, it
became necessary to consider biotransformation as a
process that affects pollutant transport and fate.
Unfortunately, many ground-water scientists adopted
the conceptual ideas most frequently used to
describe biotransformations in surface waters. In
ground water, contaminant residence time is usually
long, at least weeks or months, and frequently years
14
-------
Table 1-4 Exchange Capacities of Minerals and Rocks
Mineral
Talc
Basalt
Pumice
Tuff
Quartz
Feldspar
Kaolinite
Kaolinite (colloidal)
Nontronite
Saponite
Beidellite
Pyrophyllite
Halloysite • 2H20
Illite
Chlorite
Shales
Glauconite
Sepiolite-attapulgite-palygorskite
Diatomite
Halloysite • 4H2O
Allophane
Montmorillonite
Silica gel
Vermiculite
Zeolites
Organic substances in soil and recent
sediments
Feldspathoids
Leucite
Nosean
Sodalite
Cancrinite
For Cations
(meq/100g)
Grim (1968) Carroll (1959)
—
-
—
—
-
-
3-15
—
—
-
-
-
5-10
10-40
10-40
—
—
3-15
—
40-50
25-50
80-150
-
100-150
100-300
150-500
—
—
—
_
0.2
0.5-2.8
1.2
32.0-49.0
0.6-5.3
1.0-2.0
-
-
—
-
-
4.0
-
10-40
10-40?
10-41.0
11-20
20-30
25-54
-
~70
70-100
80
100-150
230-620
460
880
920
1,090
For Anions
(meq/100g)
Grim (1968)
—
-
—
—
-
-
6.6-13.0
20.2
12.0-20.0
21.0
21.0
-
-
-
—
—
—
—
—
-
-
23-31
-
4
—
_
—
—
_
_
Source: Matthess, 1982.
or decades. Further, contaminant concentrations that
are high enough to be of environmental concern are
often high enough to elicit adaptation of the microbial
community. For example, the U.S. EPA maximum
contaminant level (MCL) for benzene is 5 pg/l. This is
very close to the concentration of alkylbenzenes
required to elicit adaption to this class of organic
compounds in soils. As a result, the biotransformation
rate of a contaminant in the subsurface environment
is not a constant, but increases after exposure to the
contaminant in an unpredictable way. Careful field
work has shown that the transformation rate in
aquifers of typical organic contaminants, such as
alkylbenzenes, can vary as much as two orders of
magnitude over a meter vertically and a few meters
horizontally. This surprising variability in
transformation rate is not related in any simple way to
system geology or hydrology.
Biological activity may, however, promote or catalyze
chemical reactions. Stimulation of the native microbial
population and the addition of contaminant specific
"seed" microorganisms for the restoration of
contaminated aquifers by in situ biological treatment
is the subject of much current research (Canter and
Knox, 1985).
7.5.2 Contaminant Plume Behavior
The physical mechanisms of advection and
dispersion, as well as a variety of chemical and
microbial reactions, will interact to influence the
movement of contaminants in ground water. The
degree to which these mechanisms influence
15
-------
contaminant movement is dependent on a number of
factors:
a) Geologic material properties
The rate of ground-water movement is largely
dependent on the type of geologic material through
which it is moving. More rapid movement can be
expected through coarse-textured materials such as
sand or gravel than through fine-textured materials
like silt and clay. The physical and chemical
composition of the geologic material is equally
important. Fine-grained materials present the most
favorable environment for potential retardation.
b) Hydrogen ion activity (pH)
The pH of the geologic materials and the waste
stream may be a major influence on actual
retardation. The pH affects the speciation of many
dissolved chemical constituents which determine
solubility and reactivity. Ion exchange and hydrolysis
reactions are also particularly sensitive to pH.
c) Leachate composition
The influence all other factors will have on
contaminant migration ultimately depends on the
composition of the leachate or contaminants entering
the ground-water system. Similar contaminants may
behave differently in the same environment due to the
influence of other constituents in a complex leachate.
Solubility (which affects the mobile concentration),
density, chemical structure, and many other
properties can affect net contaminant migration. For
example, Figure 1-17 illustrates the appearance of
two chemicals, benzene and chloride, in a monitoring
well. Even though both contaminants may have
entered the ground-water system at the same time
and concentration, their detection in the monitoring
well reveals significantly different migration rates.
Chloride has migrated essentially unaffected while
benzene has been retarded significantly. This type of
relationship can be reversed if there is a solvent
phase in the aquifer.
Sources releasing a variety of contaminants will
create complex plumes composed of different
constituents at downgradient positions. An idealized
plume configuration composed of five different
contaminants (A-E) moving at different rates through
the ground-water system is shown in Figure 1-18.
Because of this, "the onset of contamination at a
supply well may mark the front of a set of overlapping
plumes of different compounds advancing at different
rates, which may affect the well in sequence for
decades even if the original contaminant source is
removed" (Mackay ef a/., 1985).
The effect of contaminant density on transport in
ground-water systems is presented in Figure 1-19.
Substances with densities less than water may "float"
on the surface of the saturated zone. Similarly,
substances with densities greater than water can sink
through the saturated zone until an impermeable layer
is encountered. In the situation shown in Figure 1-
19, the surface of an underlying, impermeable
formation slopes opposite to the direction of ground-
water flow in the overlying formation. Dense
contaminant movement will follow the slope of the
impermeable boundary while some dissolved product
will move with the ground water.
d) Source characteristics
Source characteristics include source mechanism
(i.e., infiltration, direct migration, interaquifer
exchange, ground-water/surface water interaction),
type of source (particularly point or nonpoint
origination), and temporal features. Aspects of source
mechanism and type have been discussed earlier.
The manner in which a contaminant is released over
time and the time which has elapsed since the
contaminant was released will greatly affect the extent
and configuration of the contaminated zone.
Figure 1-20 presents the effects caused by changes
in the rate of waste discharge on plume size and
shape. In the first case, plume enlargement results
from an increase in the rate of waste discharge to the
ground-water system. Similar effects can be
Figure 1-17
Initial
Benzene and chloride appearance in a
monitoring well (from Geraghty and Miller,
1985).
Well
Chloride
and
Benzene
Distance -
Some Time Later
Well
Distance -
16
-------
Figure 1-18 Constant release but variable constituent
source (from LeGreud, 1965).
Waste Site
Downstream Limit
of Contaminants
produced if the retardation capacity of the geologic
materials is exceeded or if the water table rises closer
to the source causing an increase in dissolved
constituent concentration. Decreases in waste
discharge, lowering of the water table, retardation
through sorption, and reductions in ground-water
flow rate can diminish the size of the plume. Stable
plume configurations suggest that the rate of waste
discharge is at steady state with respect to
retardation and transformation processes. A plume
will shrink in size when contaminants are no longer
released to the ground-water system and a
mechanism to reduce contaminant concentrations is
present. Unfortunately, many contaminants,
particularly complex chlorinated hydrocarbons and
heavy metals, may persist in ground water for
extremely long time periods without appreciable
transformation. Lastly, an intermittent or seasonal
source can produce a series of plumes which are
separated by the advection of ground water during
periods of no contaminant discharge.
1.6 Summary
To properly assess and predict the effect of ground-
water contamination at a given site, detailed
information about the nature of the suspected
contaminants, the volume of contaminants disposed
and released, the time period over which
contaminants were released, and the areas in which
contaminants were released is needed. For complex
sites, such as industrial facilities and hazardous waste
disposal landfills, this information may be limited. The
transport and fate of contaminants in ground water
must also be considered; these are often affected by
a site-specific interrelationship of physical, chemical,
biological, and temporal processes. Knowledge of site
geology, hydrology, source characteristics and
mechanisms must also precede an intelligent
investigation of ground-water contamination.
1.7 References
Broadbent, F.E., and J.B. Ott. 1957. Soil Organic
Matter - Metal Complexes: I. Factors Affecting
Various Cations. Soil Science 83:419-427.
Brown, M. 1979. Laying Waste, The Poisoning of
America by Toxic Chemicals. Pantheon Books, New
York, NY.
Burmaster, D.E., and R.H. Harris. 1982. Groundwater
Contamination: An Emerging Threat. Technology
Review 85(5):50-62.
Canter, L.W., and R.C. Knox. 1985. Ground Water
Pollution Control. Lewis Publishers, Inc. Chelsea, Ml.
Coniglio, W. 1982. Criteria and Standards Division
Briefing on Occurrence/Exposure to Volatile Organic
Chemicals. U.S. Environmental Protection Agency,
Office of Drinking Water, Cincinnati, OH.
Deutsch, M. 1961. Incidents of Chromium
Contamination of Ground Water in Michigan.
Proceedings of 1961 Symposium, Ground Water
Contamination, U.S. Department of Health, Education,
and Welfare, April 5-7, 1961, Cincinnati, OH.
Duguid, J.O. 1975. Status Report on Radioactivity
Movement from Burial Grounds in Melton and Bethel
Valleys. Environmental Science Publication No. 658,
ORNL-5017, Oak Ridge National Laboratory, Oak
Ridge, TN.
Epstein, S.S., L.O. Brown, and C. Pope. 1982.
Hazardous Waste in America. Sierra Club Books, San
Francisco, CA.
Epstein , S.S. 1979. The Politics of Cancer. Anchor
Press/Doubleday. Garden City, NY.
Fetter, C.W., Jr. 1980. Applied Hydrogeology. Charles
E. Merrill Publishing Company, Columbus, OH.
Freeze, R.A., and J.A. Cherry. 1979. Groundwater.
Prentice-Hall, Inc., Englewood Cliffs, NJ.
17
-------
Figure 1-19 Effects of density on migration of contaminants (from Geraghty and Miller, 1985).
Source of Product
(Greater Density Than Water)
Source of Product
(Lesser Density Than Water)
Unsaturated
Zone
- -* _ WaterJTable
Product Flow
Direction of
Ground-Water Flow
18
-------
Figure 1-20 Changes in plumes and factors causing the changes (modified from U.S. EPA, 1977).
A
r >
v
Enlarging
Plume
1. Increase in rate of
discharged wastes
2. Sorption activity
used up
3. Effects of changes in
water table
Reducing
Plume
1. Reduction in wastes
2. Effects of changes in
water table
3. More effective
sorption
4. More effective
dilution
5. Slower movement
and more time for
decay
Contaminated zone
Former boundary
Present boundary
Waste site
Nearly Stable
Plume
1. Essentially same
waste input
2. Sorption capacity
not fully utilized
3. Dilution effect fairly
stable
4. Slight water-table
fluctuation or effects
of water-table
fluctuation not
important
Shrunken
Plume
Waste no longer
disposed and no
longer leached at
abandoned waste
site
/ i
I '
0
Series of
Plumes
Intermittent or
seasonal source
Gass, I.E. 1980. To What Extent Is Ground Water
Contaminated? Water Well Journal 34(11):26-27.
Geraghty, J.J., and D.W. Miller. 1985. Fundamentals
of Ground Water Contam ination, Short Course
Notes. Geraghty and Miller, Inc., Syosset, NY.
Griffin, R.A., and S.F.J. Chou. 1980. Attenuation of
Polybrominated Biphenyls and Hexachlorobenzene by
Earth Materials. Environmental Geology Notes 87,
Illinois State Geological Survey, Urbana, IL.
Griffin, R.A., K. Cartwright, N.F. Shimp, J.D. Steele,
R.R. Buch, W.A. White, G.M. Hughes, and R.H.
Gilkeson. 1976. Alteration of Pollutants in Municipal
Landfill Leachate by Clay Minerals: Part I. Column
Leaching and Field Verification. Illinois State
Geological Survey Bulletin 78, Illinois State Geological
Survey, Urbana, IL.
Hughes, G.M., R.A. Landon, and R.N. Farvolden.
1971. Hydrogeology of Solid Waste Disposal Sites in
Northeastern Illinois. Solid Waste Management
Series, Report SW-124, U.S. Environmental
Protection Agency.
Illinois Environmental Protection Agency. 1986. A
Plan for Protecting Illinois Groundwater. Illinois
Environmental Protection Agency, Springfield, IL.
Killey, R.W., J.O. McHugh, D.R. Champ, E.L. Cooper,
and J.L. Young. 1984. Subsurface Cobalt-60
Migration from a Low-Level Waste Disposal Site.
Environmental Science and Technology 18(3): 148-
156.
Leckie, J.O., J.G. Pace, and C. Halvadakis. 1975.
Accelerated Refuse Stabilization Through Controlled
Moisture Application. Unpublished report, Department
of Environmental Engineering, Stanford University,
Stanford, CA.
LeGrand, H.E. 1965. Patterns of Contaminated Zones
of Water in the Ground. Water Resources Research
1(1):83-95.
Lehr, J.H. 1982. How Much Ground Water Have We
Really Polluted? Ground Water Monitoring Review
2(1 ):4.
Mackay, D.M., P.V. Roberts, and J.A. Cherry. 1985.
Transport of Organic Contaminants in Groundwater
Environmental Science and Technology. 19(5):384-
392.
Magnuson, Ed. 1980. The Poisoning of America.
Time (7):58.
Mallmann, W.L., and W.N. Mack. 1961. Biological
Contamination of Ground Water. Proceedings of 1961
Symposium, Ground Water Contamination, U.S.
19
-------
Department of Health, Education, and Welfare, April
5-7, 1961, Cincinnati, OH.
Marrin, D.L. 1985. Delineation of Gasoline
Hydrocarbons in Groundwater by Soil Gas Analysis.
Tracer Research Corporation, Tucson, AZ.
Matthess, G. 1982. Die Beschaffenheit des
Grundwassers (The Properties of Groundwater). John
Wiley and Sons, New York, NY.
Michels, Nabert, Udluft, and Zimmerman. 1959.
Expert Opinion on Questions of Protection of Aquifers
Against Contamination of Ground Water.
Bundesministerium fur Atomkernenergie and
Wasserwirtschaft, Bad Godesberg.
Middleton, M., and G. Walton. 1961. Organic
Chemical Contamination of Ground Water.
Proceedings of 1961 Symposium, Ground Water
Contamination, U.S. Department of Health, Education,
and Welfare, April 5-7, 1961, Cincinnati, OH.
Miller, D.W., ed. 1980. Waste Disposal Effects on
Ground Water. Premier Press. Berkeley, California.
512pp.
O'Hearn, M., and S.C. Schock. 1984. Design of a
Statewide Ground-Water Monitoring Network for
Illinois. Illinois State Water Survey Contract Report
354, Illinois State Water Survey, Champaign, IL.
Prescott, S.C., and M.P. Horwood. 1935. Sedgwick's
Principles of Sanitary Science and Public Health. The
MacMillan Company, New York, NY.
Rovers, F.A., H. Mooij, and G.J. Farquhar. 1976.
Contaminant Attenuation - Dispersed Soil Studies.
In: Residual Management by Land Disposal, edited by
W.H. Fuller. EPA-600/9-76-015, U.S.
Environmental Protection Agency, Cincinnati, OH.
Schiffman, A. 1985. The Trouble With RCRA. Ground
Water 23(6):726-734.
Schwarzenbach, R., W. Giger, E. Hoehn, and J.
Schneider. 1983. Behavior of Organic Compounds
During Infiltration of River Water to Ground Water -
Field Studies. Environmental Science and Technology
17(8):472-479.
Senkan, S.M., and N.W. Stauffer. 1981. What To Do
With Hazardous Waste? Technology Review
(11/12):34-37.
U.S. Congress. 1984. Protecting the Nation's
Groundwater from Contamination, Vols. I and II.
OTA-0-233 and OTA-0-276, Office of
Technology Assessment, U.S. Government Printing
Office, Washington, DC.
U.S. Department of Health, Education, and Welfare.
1961. Proceedings of 1961 Symposium, Ground
Water Contamination, April 5-7, 1961, Cincinnati,
OH.
U.S. Environmental Protection Agency. 1984. National
Primary Drinking Water Regulations, Volatile Organic
Chemicals. Federal Register, 49:24331-24355.
U.S. Environmental Protection Agency. 1983. Process
Design Manual: Land Application of Municipal Sludge.
EPA-625/1-83-016, U.S. Environmental Protection
Agency, Municipal Environmental Research Lab,
Cincinnati, OH.
U.S. Environmental Protection Agency. 1979.
Environmental Assessment: Short-Term Tests for
Carcinogens, Mutagens and Other Genotoxic Agents.
Health Effects Research Laboratory, Research
Triangle Park, NC.
U.S. Environmental Protection Agency. 1979. A Guide
to the Underground Injection Program.
Westrick, J.J., J.W. Mello, and R.F. Thomas. 1983.
The Ground Water Supply Survey: Summary of
Volatile Organic Contaminant Occurrence Data. U.S.
Environmental Protection Agency, Office of Drinking
Water, Cincinnati, OH.
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CHAPTER 2
GROUND-WATER QUALITY INVESTIGATIONS
Within the last decade, a substantial number of
ground-water quality investigations have been
conducted. Most of these have centered on specific
sites, sites that by one means or another were known
or suspected to be contaminated. In general, the sites
covered only several acres or a few square miles.
The cost of these investigations usually has been
excesssive, largely because of analytical costs. The
most disconcerting feature of many of them, however,
is that to one degree or another they were found to
be inadequate. This, in turn, necessitated additional
work and expense in response to the ever present
desire for additional information. It should be
recognized that the data base will always be
inadequate, and eventually there will be a finite sum
that is dictated by time, common sense, and
budgetary constraints. One simply has to do the best
one can with what is available.
It is suspected that the major reason that many field
investigations are both inadequate and expensive is
that a comprehensive experimental work plan was not
formulated before the project was initiated or if it was,
then it probably was not followed. Any type of
investigation must be carefully planned, keeping in
mind the overall purpose, time limitations, and project
funding. Moreover, the plan must be based on sound,
fundamental principles and a practical approach. As
far as ground-water quality investigations are
concerned, the basic questions are (1) Is there a
problem? (2) Where is it? and (3) How severe is it? A
subsequent question may relate to what can be done
to reduce the severity of the problem, that is, aquifer
restoration.
2.1 Types of Ground-Water Quality
Investigations
Ground-water quality investigations can be divided
into three general types: regional, local, and site
evaluations. The first, which may encompass several
hundred or even thousands of square miles, is
reconnaissance in nature, and is used to obtain an
overall evaluation of the ground-water situation. A
local investigation is conducted in the vicinity of a
contaminated site, may cover a few tens or hundreds
of square miles, and is used to determine local
ground-water conditions. The purpose of the site
evaluation is to ascertain, with a considerable degree
of certainty, the extent of contamination, its source or
sources, hydraulic properties, and velocity, as well as
all of the other related controls on contaminant
migration.
2.1.1 Regional Investigations
This broad brush type of investigation, which is
reconnaissance in nature, can be the starting point for
two general types of explorations. First, it can be
carried out with the purpose of locating potential
sources or sites of ground-water contamination.
Second, it will provide an understanding of the
occurrence and availability of ground water on a
regional scale. The underlying objectives are first, to
determine if a problem exists, and second, if
necessary, to ascertain prevalent hydrologic
properties of earth materials, generalized flow
directions in both major and minor aquifers, the
primary sources and rates of recharge and discharge,
the chemical quality of the aquifers and surface
water, and locations and yields of pumping centers.
These data can be useful in more detailed studies
because they provide information on the geology and
flow direction, both of which impact the local situation.
2.7.2 Local Investigations
Investigations of this nature usually include a few
square miles. The purpose is to define in greater
detail the geology and hydrology in an area
surrounding a specific site or sites of concern. Both
the geology and hydrology are likely to exert some
control on contaminant migration and nearby rock
units may be impacted as well.
2.7.3 Site Investigations
The site investigation is the most detailed, complex,
costly, and, from a legal and restoration viewpoint, the
most critical of the three types of evaluations. This
examination must address the local controls on
contaminant migration, including the geology, soil,
microbiology, geochemical interactions, and mass
flow rate of contaminant to the water table, among
others. At the same time, auxiliary investigations at
the site might include tank inventories, toxicological
evaluations, air pollution monitoring, manufacturing
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procedures, and manifest scrutiny, as well as many
other studies, all of which will eventually interface in
the development of a comprehensive report.
2.2 Conducting the Investigation
Regardless of the complexity or detail of the
investigation, a logical series of steps should be
followed. Of course, each inquiry is unique but,
nonetheless, the general rules prevail. The steps are
as follows:
1) Establish the objectives of the study.
2) Collect data.
3) Compile data.
4) Interpret data.
5) Develop conclusions.
6) Present results.
2.2.1 Establish the Objectives of the Study
Establishing the major objective of the study is
paramount. The approach, time requirements, and
funding can be vastly different between a regional
reconnaissance evaluation and a site investigation.
The former, which deals with gross features, may
require only days while the latter, which necessitates
minute detail, may demand years. The statement of
the objective can be as simple as "Develop a general
understanding of the regional ground-water
situation" to "Evaluate the degradation and dispersion
of selected organic compounds in the capillary zone
at the A site."
In both of the above examples the objective is clearly
stated and the complexity is evident.
Once the general objective is established, a number
of secondary purposes must be considered. These
involve the physical system and the chemical aspect.
Secondary objectives include the following:
1) Determination of the thickness, soil
characteristics, infiltration rate, and water-
bearing properties of the unsaturated zone.
2) Determination of the geologic and hydrologic
properties and dimensions of each unit in the
geologic column that potentially could be
impacted by ground-water contamination. This
includes rock type, thickness of aquifers and
confining units, areal distribution, structural
configuration, transmissivity, hydraulic
conductivity, storativity, water levels, infiltration or
leakage rate, and rate of evapotranspiration, if
appropriate.
3) Determination of recharge and discharge areas, if
appropriate.
4) Determination of the direction and rate of
ground-water movement in potentially impacted
units.
5) Determination of the ground water and surface
water relationships.
6) Determination of the background water quality
characteristics of potentially impacted units.
7) Determination of potential sources of
contamination and types of contaminants.
2.2.2 Data Collection
Data collection forms the basis for the entire
investigation and, consequently, time must be
expended and care exercised in carrying out this task.
The amount and types of data to be collected are
dictated by the objectives of the study. Before the
field is ever visited, a thorough search should be
made of files and the literature for pertinent
information. Materials that should be collected, if
available, include soil, geologic, topographic, county,
and state maps, geologic cross sections, aerial
photographs, satellite imagery, location of pumping
centers and discharge rates, well logs, climatological
and stream dische'ge records, chemical data, and the
locations of potential sources of ground-water
contamination.
Many of these data are readily available in files or
report form and can be obtained from an assortment
of state and Federal agencies. Personnel with these
agencies also can be of great help owing to their
knowledge of the state or county and their familiarity
with the literature. Examples include the U.S.
Geological Survey, which has at least one office in
each state, the state geological survey and several
state agencies that deal with water, such as the state
water survey, water resources board, or the water
commission. Some states have several commissions
or boards that are involved with water. Other sources
of information include the state or Federal department
of agriculture, Soil Conservation Service, and weather
service, among others.
Climatological data are important because they
indicate precipitation rates and patterns, both of which
influence surface runoff, runoff, and ground-water
recharge. Additionally, climatological data include
temperature measurements, which can be used in the
evaluation of evapotranspiration. Evapotranspiration of
shallow ground water can produce a significant effect
on the water-table gradient, causing it to change in
slope and direction, not only seasonally, but diurnally
as well.
Stream discharge and chemical quality records can
be used for several types of regional and local
evaluations, as described in Chapter 4.
Soil types are related to the original rock from which
they were derived. Consequently, soil maps can be
used as an aid in geologic mapping. Soil information
is necessary also to evaluate the potential for
22
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movement of organic and inorganic compounds
through the unsaturated zone.
Exceedingly useful tools, both for office and field
study, are aerial photographs and satellite imagery.
The latter should be examined first in an attempt to
detect trends of lineaments, which may indicate the
presence of faults or joints. These may reflect zones
of high permeability that exert a strong influence on
fluid movement from the land surface or through the
subsurface. Satellite imagery also can be used to
detect the presence of shallow ground water owing to
the subtle tonal changes and differences in vegetation
brought about by the higher moisture content. Rock
types may be evident also on imagery.
Aerial photographs, particularly stereoscopic pairs,
should be an essential ingredient of any hydrologic
investigation. They are necessary to further refine the
trends of lineaments, map rock units, determine the
location of cultural features and land use, locate
potential drilling sites, and detect possible sources of
contamination. Topographic and state and county
road maps also are useful for many of these
purposes.
Geologic reports, maps, and cross sections provide
details of the surface and subsurface, including the
areal extent, thickness, composition, and structure of
rock units. It must be remembered that geology is the
key to any ground-water investigation. These
sources of information should be supplemented, if
possible, by an examination of the logs of wells and
test holes. Depending on the detail of the logs, they
may provide a clear insight into the complexities of
the subsurface.
Logs of wells and test holes are essential in ground-
water investigations. They provide first-hand
information on types and characteristics of rocks in
the subsurface, their thickness, and areal extent.
Logs also may describe drilling conditions that allow
one to infer relative permeability values (see Chapter
9), describe well construction details, and report
water-level measurements.
Chemical data may be available from reports, but the
most recent information is probably stored in local,
state, or Federal files. Concentrations of selected
constituents, such as dissolved solids, specific
conductance, chloride, and sulfate, should be plotted
on base maps and used to estimate background
quality and, perhaps, detect places of contamination.
Both surface and ground-water quality data should
be examined.
Chemical analyses that report concentrations of
organic compounds are bound to be sparse and even
those are likely to be questioned for one reason or
another. Only within the last decade or so have
organic compounds become of concern. The cost of
analysis is high, and much remains to be learned
about appropriate methods of collection, storage,
analysis, and interpretation. Consequently,
investigators, whenever possible, will need to rely on
analyses of inorganic substances to detect sites of
ground water contaminated by these complex
substances. On the other hand, reliance on
concentrations of inorganic constituents to evaluate
contamination by organic compounds may not always
be appropriate, possible, or desirable. In many
situations, however, both organic and inorganic
substances are present in a leachate.
2.2.3 Field Investigation
Once an exhaustive search of the literature, files,
maps, aerial photography, and satellite imagery has
been conducted and, at least to some extent, relevant
information has been studied, it is appropriate to visit
the field. During the office evaluation, several points
should become reasonably clear. These include a
general appreciation of the regional hydrogeology,
geology, and water quality, as well as water use and
areas of potential problems.
These characteristics should be verified during field
examinations. Initially field work should be
reconnaissance in nature. The complexity or detail of
the field work will expand with time in response to
increasing familiarity with the region, area, or site, as
well as the objectives of the study.
During early field visits, particular attention should be
paid to landforms, streams and stream patterns,
locations of springs, seeps, and lakes, as well as
vegetation. Landforms are controlled by the geology
and many hills are capped by resistant strata, such as
sandstone, while valleys are usually carved into soft,
less resistant material, such as shale. Likewise, many
changes in topographic slope are related to
differences in rock type. These, in turn, provide a
general impression of the types of rocks present, their
areal extent, and composition. Rock exposures in
stream channels and road cuts are very useful also
when attempting to understand the local geology.
Joint and fracture systems, their directional trends,
density, and size can all be measured on rock
outcrops. Fluid movement through joints and other
fractures may control entirely, or nearly so, the
migration of contaminants.
Stream patterns also are related to the geology,
especially geologic structure and fracture or joint
systems. A brief examination of streams in the region
is useful since it provides an idea of the relative
difference in discharge from one stream to another.
As indicated in Chapter 4, streams can provide a
wealth of information on basin permeability, shallow
ground-water quality, and local sites where the
ground water is contaminated.
Springs and seeps are zones of ground-water
discharge. They should exist in the vicinity of strata of
low permeability that are overlain by a unit of greater
permeability, that is, an aquitard overlain by an
23
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aquifer. Rarely does water continually discharge from
springs and seeps. Most commonly the discharge is
greater during spring and early summer, during the
fall rainy period, or during and after a period of
precipitation. Following these intervals of ground-
water recharge, the discharge of springs and seeps
diminishes or ceases entirely as the water-bearing
zone becomes depleted. Nonetheless, the area
downslope from the discharge zone has a higher
moisture content and commonly supports far more
vegetation, both grasses and trees, than is present in
adjacent areas. The presence of the vegetation may
allow the mapping of certain rock types.
2.3 Regional Investigations
Regional investigations are conducted for many
different purposes. One type is to detect potential
sources and locations of ground-water
contamination. An example was described in Chapter
4 in which surface water data were used to detect
potential sources of contamination (abandoned and
producing oil wells and salt-water disposal ponds)
and locate relatively small areas in which the ground
water was contaminated by these activities in Alum
Creek basin in central Ohio.
Another type of exceedingly broad scope includes
library searches. Examples include an early EPA
effort to evaluate ground-water contamination
throughout the United States (van der Leeden et a/.,
1975; Miller and Hackenberry, 1977; Scalf ef a/.,
1973; Miller et at., 1974; and Fuhriman and Barton,
1971). The reports are useful for obtaining a general
appreciation of the major sources of contamination
and the magnitude over a regionally extensive area.
An excellent description of the geology and hydrology
of the Ohio River basin was prepared by Deutsch et
a/. (1969). The 10 volume manuscript depicted each
subbasin in considerable though broad detail. It
served as the basis for a subsequent report that
related ground-water quality and streamflow
throughout the Ohio River basin. The report was
prepared for the Federal Water Pollution Control
Agency and, although completed in 1968, the report,
unfortunately, was never published; draft copies
should be available in EPA files. Examination of any
of the volumes of either of the reports would provide
an investigator with many ideas on how to conduct a
regional evaluation.
In 1980 individuals in EPA Region VII became aware
of what appeared to be a large number of wells that
contained excessive concentrations of nitrate.
Suspecting a widespread problem, a regional
reconnaissance investigation was initiated. The
general approach consisted of a literature search, a
meeting in each state with regulatory and health
personnel, an evaluation of existing data, and an
interpretation of all of the input values.
The fundamental principle guiding this study was the
fact that abnormal concentrations of nitrate can arise
in a variety of ways, both from natural and manmade
sources or activities. The degradation may
encompass a large area if it results from the
overapplication of fertilizer and irrigation water on a
coarse textured soil, from land treatment of
wastewaters, or from a change in land use, such as
converting grasslands to irrigated plots. On the other
hand, it may be a local problem affecting only a single
well if the contamination is the result of animal
feedlots, municipal and industrial waste treatment
facilities, or improper well construction/maintenance.
For the most part the data base for this study was
obtained from STORET. First, nitrate concentrations
in well waters were placed in a separate computer
file. Two maps were generated from the file, the first
showing the density of wells that had been sampled
for nitrate, and the second showing the density of
wells that exceeded 10 mg/l of nitrate (Figure 2-1).
The maps were produced by the STORET routine,
Multiple Station Plot. These maps indicated the areas
of the most significant nitrate problems. In turn, the
nitrate distribution maps were compared to geologic
maps, which allowed some general identification of
the physical system that was or appeared to be
impacted (Figure 2-2).
Iowa, eastern Nebraska, northeastern Kansas, and
the northern third of Missouri are characterized by
glacial till interbedded with local deposits of outwash.
Throughout the area are extensive deposits of
alluvium. Many of the aquifers are shallow and wells
are commonly dug, bored, or jetted. This area
contained the greatest number of domestic wells with
high nitrate concentrations. It also contained the
Figure 2-1 Location of wells with nitrate exceeding 10
mg/l in Region 7.
l>10mg/l N03
24
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Figure 2-2 Generalized rock types with high nitrate
concentrations in Region 7.
Loess
Sand
-\ V
Glacial Till
Stream Alluvium
Gravel '
l
i
Stream
\ Alluvium
greatest number of municipal wells that exceeded the
nitrate MCL (Maximum Contaminant Level). The
cause of contamination in the shallow domestic wells
was suspected to be poor well construction and
maintenance, but this was possibly not the case for
many of the generally deeper municipal wells, where
the origin appeared to be from naturally occurring
sources in the glacial till.
Most of Nebraska and western Kansas are mantled
by sand, gravel, and silt, which allow rapid infiltration.
The water table is relatively shallow. The irrigated part
of this region, particularly adjacent to the Platte River
and in areas of Holt County, Nebraska, contained the
greatest regional nitrate concentrations in the four
state area. This was brought about by the excessive
application of fertilizers and irrigation waters in this
very permeable area.
The remaining area in Kansas and an adjacent part of
Missouri is underlain by sedimentary rocks across
which flow many streams and rivers with extensive
flood plains. Most of the contaminated wells tapped
alluvial deposits. The primary cause of high nitrate in
domestic wells was suspected to be poor well
construction/maintenance or poor siting with respect
to feedlots, barnyards, and septic tanks.
The southern part of Missouri is represented by
carbonate rocks containing solution openings.
Aquifers in these rocks are especially susceptible to
contamination and the contaminants can be
transmitted great distances with practically no change
in chemistry other than dilution. The carbonate terrain
is not easily manageable, nor is monitoring a simple
technique because of the vast number of possible
entry sites whereby contaminants can enter the
subsurface.
The STORET file was also used to generate a
number of graphs of nitrate concentration versus time
for all of the wells that were represented by multiple
samples. The graphs clearly showed that the nitrate
concentration in the majority of wells ranged within
wide limits from one sampling period to the next,
suggesting leaching of nitrate during rainy periods
from the unsaturated zone.
The state seminars were exceedingly useful because
the personnel representing a number of both state
and Federal agencies had a good working knowledge
of the geology, water quality, and land-use activities
of their respective states.
Although the study extended over several months, the
actual time expended amounted to only a few days.
The conclusions, for the most part, were
straightforward and, in some cases, pointed out
avenues for improvement in sample collection and
data storage/access. The major conclusions are as
follows:
1) High levels of nitrate in ground water appear to be
randomly distributed through the region.
2) The most common cause of high nitrate
concentration in well water appears to be related
to inadquate well construction, maintenance, and
siting. Adequate well construction codes could
solve this problem. Dug wells, those improperly
sealed, and wells that lie within an obvious source
of contamination, such as a pig lot, should
probably be abandoned and plugged.
3) In areas of extensive irrigation where excess
water and fertilizer are applied to coarse textured
soils, the nitrate concentration in ground water
appears to be increasing.
4) In the western part of the region, changes in land
use, particularly the cultivation or irrigation of
grasslands, has resulted in leaching of substantial
amounts of naturally occurring nitrate from the
unsaturated zone.
5) The population that is consuming high nitrate
water supplies is small, accounting for less than 2
percent of the population.
6) There have been no more than two reported
cases of methemoglobinemia in the entire region
within the past 15 years despite the apparent
increase in nitrate concentration in ground-water
supplies. This implies a limited health hazard.
7) State agency personnel are convinced that they
do not have significant nitrate-related health
problems.
8) Many of the wells that are used in the state and
Federal monitoring networks are of questionable
value because little or nothing is known about
their construction.
9) The volume of chemical data presently in the files
of most of the state agencies within the region is
25
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not adequately represented in the STORE! data
system.
This cursory examination provided only a general
impression of the occurrence, source, and cause of
abnormal nitrate concentrations in ground water in the
region. Nonetheless, it furnished a base for planning
local or site investigations, was prepared quickly, and
did not require field work or extensive data collection.
As mentioned previously, the source of excessive
nitrate in many municipals could not be readily
explained. There could be multiple sources related to
naturally occurring high nitrate concentrations in the
unsaturated zone or the glacial till, to contamination,
or to poor well construction. Definitive answers would
require more detailed local or site studies. The overall
effect of changing from grazing land to irrigated
agriculture, in view of the great mass of nitrate-
bearing substances in the unsaturated zone that
would be leached, clearly warrants additional local
investigation. Although the concentration of nitrate in
underlying ground water would increase following
irrigation, it is likely that some control on the rate of
leaching could be implemented by limiting the amount
of water applied to the fields.
The obvious relationship between the application of
excessive amounts of fertilizer and water on a coarse
textured soil in Nebraska shows the need for
experimental work on irrigation techniques in order to
reduce the loading. Also implied is the necessity for
developing educational materials and seminars, in
order to offer means whereby irrigators can reduce
water, pesticide, and fertilizer applications and yet
maintain a high yield.
2.4 Local Investigations
Local investigations can be as varied in scope and
area! extent as regional evaluations and the difference
between the two is relative. For example, one might
desire to obtain some knowledge of the hydrogeology
of an area encompassing a few tens or several
hundred square miles in order to evaluate the effect
of oil-field brine production and disposal. Examples
of this scope include Kaufmann (1978) and Oklahoma
Water Resources Board (1975). The other extreme
may center around a single contaminated well. In this
case the local investigation would most likely focus on
the area influenced by the cone of depression, the
size of which depends on the geology, hydraulic
properties, and well discharge. Consider an area in
the Great Plains where a number of small
municipalities have reported that some of their wells
tend to increase in chloride content over a period of
months to years. The increase in a few wells has
been sufficient to cause abandonment of one or more
wells in the field. Additionally, a number of wells when
drilled yielded brackish or salty water necessitating
additional drilling elsewhere. This is an expensive
process that strains the operating budget of a small
community.
In this case, the local investigation covered an area of
576 square miles, that is, 16 townships. A review of
files and reports and discussions with municipal
officials and state and Federal regulatory agencies
indicated that the entire area had produced oil and
gas for more than 30 years. Inadequate brine disposal
was the most likely cause of the chloride problem.
During the initial stage of the investigation, all files
dealing with the quality of municipal well water were
examined and this task was followed by a review of
the geology, which included a review of all existing
maps, cross sections, and well logs, both lithologic
and geophysical.
The chemical data clearly showed that the chloride
content in some wells increased with time, although
not linearly. The geologic phase of the study showed
that the rocks consist largely of interbedded layers of
shale and sandstone and that the sandstone deposits,
which serve as the major aquifers, are lenticular and
range from 12 to about 100 feet in thickness. The
sandstones are fine-grained and cemented to some
degree and, as a result, each unit will not yield a
large supply. Resultingly, all sandstone bodies are
screened.
Trending north-south through the east-central part
of the area is an anticline (Figure 2-3) that causes
the rocks to dip about 50 feet per mile either to the
east or west of the strike of the structure (Figure 2-
4). This means that a particular sandstone will lie at
greater depths with increasing distances from the axis
of the anticline. (Refer to Chapter 9).
In this example, the subsurface geology was based
on an evaluation of geophysical and geologists' logs
of wells and test holes, including oil and gas wells
and tests. As shown in Figure 2-4, interpretation of
the logs in the form of a geologic cross section brings
to light an abundance of interesting facts. The
municipal wells range in depth from 400 to 900 feet,
but greater depth does not necessarily indicate a
larger yield nor does depth imply a particular chemical
quality. The difference in well depth and yield is
related to the thickness and permeability of the
sandstone units encountered within the well bore.
Secondly, the volume of the sandstone components
ranges widely, but the thinnest and most
discontinuous units increase in abundance westward.
More importantly, the mineral content of the ground
water, which can be determined from geophysical
logs, increases down the dip of the sandstone, from
fresh in the outcrop area, to brackish, and finally to
salt water (Figure 2-4). Notice also that brackish and
saline water lie at increasingly shallower depths to the
west of the outcrop area.
The position and depth of a few municipal wells and
test holes are also shown on the cross section. Well
26
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Figure 2-3 Generalized geologic map of a local investigation.
Scale (miles)
. Sandstone outcrop area
• Aquifer thickness exceeds 125 ft
1 would be expected to have a small yield of brackish
water. Well 2 is an abandoned test hole that
penetrated a thick saline zone as well as a thick
brackish water zone. In the case of Well 3, the fresh
water derived from the thin, shallower sandstones is
sufficient to dilute water derived from the more
mineralized zones. On the other hand, as the artesian
pressure in the shallow sandstones decreases with
pumping and time, an increasing amount of the well
yield might be derived from the deeper brackish layer,
causing the quality to deteriorate.
The major conclusion derived from this study is that
the most readily apparent source of high chloride
content in municipal wells, that is, inadequate oil-
field brine disposal, is not the culprit. Rather all of the
problems are related to natural conditions in the
subsurface, brought about by the downdip increase in
dissolved solids content as fresh water grades into
brackish and eventually into saline water.
Deterioration of municipal well water quality is related
to the different zones penetrated by the well and to a
decrease in artesian pressure in fresh water zones
brought about by pumping. The latter allows updip
migration of brackish or saline water to the well bore
or lateral or vertical leakage of mineralized water from
one aquifer to another, which again is the result of a
pressure decline in the fresh water zones. The
problem could be diminished by constructing future
wells eastward toward the axis of the anticline, limiting
them to those areas either within the outcrop or
where the thickness of the fresh water aquifers
comprise a total thickness that exceeds 125 feet
(Figure 2-3).
2.5 Site Investigations
Site investigations are ordinarily complex, detailed,
and expensive. Furthermore, the results and
interpretations are likely to be thoroughly questioned
in meetings, interrogatories, and in court because the
expenditure of large sums of money may be at stake.
The investigator must exercise extreme care in data
collection and interpretation. The early development
of a flexible plan of investigation is essential and it
must be based, at least in part, on guidelines
established by the Environmental Protection Agency,
such as the Technical Enforcement Guidance
Document. State regulatory agencies may have even
more stringent requirements.
The investigative plan needs to be flexible in a
practical way. For example, the position of all test
holes, borings, and monitoring wells should not be
determined in the office at the start of the
investigation. Rather, locations should be adjusted on
the basis of the information obtained as each hole is
completed. In this way, one can maximize the data
acquired from each drill site and more appropriately
locate future holes in order to develop a better
understanding of the ground-water system.
In the case of Superfund and RCRA sites, the
regulatory investigator probably will be required to
work with or at least use data collected by
consultants for the defendant. In some cases, the
defendant conducts and pays for the entire
investigation; regulatory personnel only modify the
work plan so that it meets established guidelines.
There are two points to consider in these situations.
First, the consultant is hired by the defendant and
should act in his best interest. This means that his
interpretations may be slighted toward his client and
concepts detrimental to the client are not likely to be
freely given. Second, even though the regulatory
investigator and the consultant, to some degree, are
adversaries, this does not mean that the consultant is
dishonest, ignorant, or that his ideas are incorrect. It
must always be remembered that the entire purpose
of the investigation is to determine, insofar as
possible, what has or is occurring so that effective
and efficient corrective action can be undertaken. In
the long run cooperation leads to success.
Several generalized methods have been available for
a number of years to evaluate a possible or existing
site relative to the potential for ground-water
contamination. These rating techniques are valuable,
in a qualitative sense, for the formulation of a detailed
27
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Figure 2-4 Geologic cross section showing downdip change in water quality.
- 1 2 3 4 5 6
1100 -
F = Fresh
B = Brackish
S = Salty
-300 -
investigation. The most noted is probably the
LeGrand (1983) system, which takes into account the
hydraulic conductivity, sorption, thickness of the water
table aquifer, position and gradient of the water table,
stream density, topography, and distance between a
source of contamination and a well or stream. The
LeGrand system was modified by the U.S.
Environmental Protection Agency (1983) for the
Surface Impoundment Assessment.
Fenn and others (1975) formulated a water balance
method to predict leachate generation at solid waste
disposal sites. Gibb and others (1983) devised a
technique to set up priorities for existing sites relative
to their threat to health. An environmental
contamination ranking system was contrived by the
Michigan Department of Natural Resources (1983).
On a larger scale is DRASTIC, which is a method,
based on hydrogeologic settings, to evaluate the
potential of ground-water contamination.
As an example of a ground-water quality site
investigation, consider a rather small refinery that has
been in existence for several decades. For some
regulatory reason an examination of the site is
required. The facility, which has not been in operation
for several years, includes an area of about 245
acres. The geology consists of alternating deposits of
sandstone and shale that dip slightly to the west; the
upper 20 to 30 feet of the rocks are weathered.
Potential sources of ground-water contamination
include wastewater treatment ponds, a land treatment
unit, a surface runoff collection pond, and a
considerable number of crude and product storage
tanks. Line sources of potential contaminants include
unimproved roads, railroad lines, and a small
ephemeral stream that carries surface runoff from the
plant property to a holding pond.
After considering the topography and potential
sources of contamination, the locations of 11 test
borings were established. The purpose of the holes
was to determine the subsurface geologic conditions
underlying the site. Following completion, the holes
were geophysically logged and then plugged to the
surface with a bentonite and cement slurry. The bore
hole data were used to determine drilling sites for 20
observation wells, in order to ascertain the quality of
the ground water, to establish the depth to water, and
to determine the hydraulic gradient. Eight of the
observation wells were constructed so that they could
be used later as a part of the monitoring system. Two
of the wells tapped the weathered shale, their
purpose being to monitor the water table, evaluate the
relation between precipitation and recharge, and
ascertain the potential fluctuation of water quality in
28
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the weathered material in order to determine if it
might serve as a pathway for contaminant migration
from the surface to the shallowest aquifer. (From a
technical perspective, the weathered shale and
sandstone is not an aquifer, but from a regulatory
point of view it could be considered a medium into
which a release could occur and, therefore, might fall
under RCRA guidelines.)
Regulations required that the shallowest aquifer be
monitored, which in this case was a relatively thin,
saturated sandstone. After the initial investigative
information was available, all of the findings were
used to design a ground-water monitoring system.
This plan called for an additional 12 monitoring wells.
Graphics based on all of the drilling information
(geologic and geophysical logs) included several
geologic cross sections (Figure 2-5) and maps
showing the thickness of shale overlying the aquifer
(Figure 2-6), thickness of the aquifer, and the
hydraulic gradient (Figure 2-7). The major purpose
of the first map was to show the degree of natural
protection that the shale provided to the aquifer
relative to infiltration from the surface. The aquifer
thickness map was needed for the design of
monitoring wells. The water-level gradient map was
necessary to estimate ground-water velocity and
flow direction. During the drilling phases, cores of the
aquifer and the overlying shale were obtained for
laboratory analyses of hydraulic conductivity, porosity,
specific yield, grain size, mineralogy, and general
description. Aquifer tests were conducted on 20 of
the wells.
The cross sections and maps indicate that the
sandstone dips gently eastward and nearly crops out
in a narrow band along the western margin of the
facility. Elsewhere, owing to the change in topography
and the dip of the aquifer, the sandstone is overlain
by 25 feet or more of shale; throughout nearly all of
the site the shale exceeds 50 feet in thickness.
Consequently, only one small part of the aquifer, its
outcrop and recharge area, is readily subject to
contamination.
The water-level map indicates that the hydraulic
gradient is not downdip but rather about 55 degrees
from it. It is controlled by the topography off site. The
average gradient is about 0.004 feet per foot, but
from one place to another it differs to some extent,
reflecting changes in aquifer thickness.
The topographic map shows that surface runoff from
the entire facility is funneled down to a detention
pond. The pond and the lower part of the drainage
way lie in the vicinity of the aquifer's recharge or
outcrop area.
Logs of the drill holes list specific depths in six of the
holes in which hydrocarbons were present. All were
reported in the unsaturated zone at depths of 2 to 9
feet with thicknesses ranging from a half inch to
Figure 2-5 Geologic cross section for the site investigation.
A'
Sandstone .'.'.•
Q2J Sandstone
E3 Shale
Potentiometric
Surface
Uppermost
Aquifer
nearly a foot. At these locations the shale overlying
the aquifer exceeded 55 feet in thickness.
Chemical analyses of water from the observation
wells indicated, with one exception, that the quality
was within background concentrations and no organic
compounds were present. The exception was an
observation well near the surface runoff retention
pond.
Precipitation in the area and the hydrograph of a well,
14 feet deep, in the weathered shale is shown in
Figure 2-8. These data show that the weathered
shale is shown in Figure 2-8. These data show that
the weathered material responds quickly to rainfall
events, despite the fact that laboratory values of
hydraulic conductivity were exceedingly low. This
strongly suggests that the weathered shale is quite
permeable, the permeability being related to fractures.
Therefore, from a hydraulic perspective, the
weathered shale appears to form a medium that
would allow the migration of contaminants from the
surface.
The relation between precipitation and nitrate, which
is a good tracer, in the weathered material is shown
in Figure 2-9. Annually the nitrate concentration
fluctuates between about 4 and 11 mg/l, which is
typical of an area characterized largely by grasslands.
The nitrate is of natural origin and the range in
concentration only indicates the variation in
29
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Figure 2-6 Map showing thickness of shale overlying the Figure 2-7 Potentiometric surface of the uppermost
uppermost aquifer. aquifer.
• Observation well
A Bore hole
30
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Figure 2-8 Relation between precipitation and water level.
879
1 £2?3«>4.o5z6a
— ? * o o •
r4.00
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0
_c
O
c
-2.00
.00
o.
'u
cu
i_
CL
0.00
1985
31
-------
Figure 2-9 Relation between precipitation and nitrate concentration.
4.00
0)
.c
o
c
3.00
c
O 2.00 :
-^
O
"5-
O
-5.00
0.00
234-567
1S84-
S 9 10 11 12 13 14 15 16
1985
32
-------
background. In this case as in every other,
background concentration is not a finite number, but
rather a range. The graph does not indicate a good
correlation between nitrate and rainfall, but here are a
few periods when the relationship is close.
Multiple analyses of nitrate in the sandstone aquifer
showed that nitrate ranged only between 2 and 4 mg/l
over a period of 14 months. This suggests that
substances that originate from the surface or
unsaturated zone do not impact the sandstone
aquifer. More likely they migrate laterally to points of
diffuse discharge along hillsides where the water is
lost by evapotranspiration. The hydrograph also
suggests that the water table declines rapidly in
response to evapotranspiration.
Evaluation of all of the data indicated two potential
problems -- hydrocarbons in the unsaturated zone
and ground-water contamination in the vicinity of the
surface runoff detention pond. Since the plant had
been in operation more than 50 years, the
hydrocarbons had migrated from the surface into the
weathered shale no more than 9 feet, and there was
a minimum of at least 45 feet of tight, unfractured
shale between the hydrocarbons and the shallowest
aquifer, it did not appear that the soil contamination
would present a hazard to ground water.
The existence of contaminated ground water,
however, was a problem that needed to be addressed
even though the sandstone aquifer is untapped and is
never likely to serve as a source of supply. Four
additional monitoring wells were installed
downgradient in order to determine the size of the
plume and its concentration. Corrective action called
for removal of sediment and sludge from the pond,
backfilling with clean material, a cap, and pumping to
capture the plume. The contaminated water was
treated on site with existing facilities.
2.6 Summary
Each ground-water quality investigation is unique,
although general guidelines need to be followed for all
of them. The investigator must first clearly define the
objectives of the study, for these will determine the
complexity, time element, and cost of the project.
Specific techniques that might be required are
described in other chapters in this report.
2.7 References
Deutsch, M., P. Jordan, and J. Wallace. 1969. Ohio
River Basin Comprehensive Report, Appendix E.,
Ground Water. Corps of Engineer Division, Ohio
River, Cincinnati, OH.
Fenn, D.G., KJ. Hanley, and T.V. DeGeare. 1975.
Use of the Water Balance Method for Predicting
Leachate Generation from Solid Waste Disposal
Sites. U.S. Environmental Protection Agency Solid
Waste Report No. 168, Cincinnati, OH.
Fuhriman, O.K., and J.R. Barton. 1971. Ground Water
Pollution in Arizona, California, Nevada and Utah.
16060 ERU 12/71, U.S. Environmental Protection
Agency.
Gibb, J.P., M.J. Barcelona, S.C. Schock, and M.W.
Hampton. 1983. Hazardous Waste in Ogle and
Winnebago Counties, Potential Risk Via Ground
Water Due to Past and Present Activities. Doc. No.
83/26, Illinois Department of Energy and Natural
Resources.
Kaufmann, R.F. 1978. Land and Water Use Effects on
Ground-Water Quality in Las Vegas Valley.
EPA/600/2-78/179, U.S. Environmental Protection
Agency.
LeGrand, H.E. 1983. A Standardized System for
Evaluating Waste-Disposal Sites. National Water
Well Association, Worthington, OH.
Michigan Department of Natural Resources. 1983.
Site Assessment System (SAS) for the Michigan
Priority Ranking System under the Michigan
Environmental Response Act. Michigan Department
of Natural Resources.
Miller, D.W., F.A. DeLuca, and T.L. Tessier. 1974.
Ground Water Contamination in the Northeast States.
EPA/660/2-74/056, U.S. Environmental Protection
Agency.
Miller, J.C. and P.S. Hackenberry. 1977. Ground-
Water Pollution Problems in the Southeastern United
States. EPA/600/3-77/012, U.S. Environmental
Protection Agency.
Oklahoma Water Resources Board. 1975. Salt Water
Detection in the Cimarron Terrace, Oklahoma.
EPA/660/3-74/033, U.S. Environmental Protection
Agency.
Scalf, M.R., J.W. Keeley, and C.J. LaFevers. 1973.
Ground Water Pollution in the South Central States.
EPA/R2-73/268, U.S. Environmental Protection
Agency.
U.S. Environmental Protection Agency. 1983. Surface
Impoundment Assessment National Report.
EPA/570/9-84/002, U.S. Environmental Protection
Agency.
33
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van der Leeden, F., L.A. Cerrillo, and D.W. Miller.
1975. Ground-water Pollution Problems in the
Northwestern United States. EPA/660/3-75/018, U.S.
Environmental Protection Agency.
34
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CHAPTERS
GROUND-WATER RESTORATION
A number of techniques are available to either contain
a pollutant and/or treat the ground water and at least
partially clean up a contaminated aquifer. These
techniques range from removal of the polluted
material and physical, chemical, or biological
treatment on the surface, to physical containment and
in-situ treatment with chemicals or microbes. Most
of the available technologies have been developed
through remedial activities in the Superfund program.
The major emphasis of this chapter will be an
overview of the remedial and restoration technology
which can be considered for application in aquifer
clean-up operations, with special emphasis on
ground-water pumping systems and in-situ
bioreclamation. Most of the discussions will concern
hydrocarbons because there is more information on
this particular contaminant. A number of the newer
technologies, such as various in-situ biodegradation
techniques, where applicable, are indicated as
potentially very cost effective. The most significant
benefit of in-situ treatment technologies is that
physical removal of contaminated soils and pollutants
is eliminated; this significantly reduces cost and
public health risk.
3.1 Subsurface Effects on Contaminant
Mobility
The movement of most ground-water contaminants
is controlled by gravity, the permeability and wetness
of the geological materials receiving them, and the
miscible character of the contaminants in ground
water. When material, particularly a hydrocarbon, is
released to the soil, it is actively drawn into the soil
by capillary attraction and by gravity. As the main
body of materials moves down into the moister
regions of the soil, the capillarity becomes less
important and the materials move through the most
favorable channels by displacing air, eventually
reaching the water table where components less
dense than water spread laterally along the air-water
interface. In instances involving a heavier
contaminant, the material continues to move
downward in the saturated zone. In both cases the
contaminants migrate downgradient with the natural
ground-water flow (Wilson and Conrad, 1984).
The quantity of a contaminant such as hydrocarbons
that reaches the water table is dependent both on the
quantity involved and the nature of the earth
materials. The coarser the earth materials, the larger
the amount that will reach the ground water. The
entire volume of hydrocarbon may be immobilized in
the unsaturated zone, although it may continue to
migrate downgradient where it becomes a threat to
the quality of ground water. Material immobilized in
the vadose zone may remain there unless it is
physically, chemically, or biologically removed.
The hydrocarbon liquid phase is generally referred to
as being immiscible with both water and air. However,
it is important to realize that various components of
the hydrocarbon volatilize into the air phase and
dissolve into the water phase. A halo of dissolved
hydrocarbon components precedes the immiscible
phase, some of which is trapped in the pore space
and left behind. The trapped hydrocarbon remains as
pendular rings and/or isolated immobile blobs. Even
when the so-called residual immiscible hydrocarbon
is exhausted by immobility, ground water coming into
contact with the trapped material leaches soluble
hydrocarbon components and continues to
contaminate ground water.
With two fluid phases, water and hydrocarbon, the
residual hydrocarbon is trapped by one or both of two
mechanisms known as by-passing and snap-off.
Snap-off depends strongly on pore shape and
wetability of the soil particles. In high aspect ratio
pores, in which the pore throats are much smaller
than pore bodies, snap-off is common. Snap-off
occurs as water moving through the small throats
begins to go around the outer surface of larger
droplets, thereby isolating them in the larger voids.
By-passing occurs as hydrocarbons are routed
around soil grains where the branched pore canals
are of unequal size. The velocity of water passing
through the smaller diameter branched pore is faster
and therefore travels around the grain faster in one
branch than the other. This results in trapping of
hydrocarbon in the larger slower moving pore canal
(Wilson and Conrad, 1984).
Residual hydrocarbon can occupy from 15 to 40
percent or more of the pore space as a result of
35
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these trapping processes depending on several
physical characteristics of the subsurface. The ability
to design and conduct successful remediation
strategies depends in large part on the ability to
understand, predict and enhance the mobility of both
liquid and dissolved hydrocarbons. Theoretically,
there are a number of ways the trapped residuals can
be mobilized and caused to concentrate where much
lower content of residual product will have to be dealt
with. Obviously the most used mobilization technique
is to increase the hydraulic gradient, usually by
pumping, thereby increasing the Darcy Velocity of the
water phase in the saturated zone. When the velocity
is increased sufficiently some of the blobs begin to
move. A critical element in this mobilization process
is the length of the blob in the direction of flow. The
gradient must be high enough to squeeze the blobs
through pore throats. After they are mobilized, blobs
do not maintain their size and shape. The large
mobilized blobs break up into smaller blobs with a
significant fraction being only temporarily mobilized.
Blobs may also coalesce and become trapped at
greater distances from the source. Active research in
physical mobilization technology is progressing rapidly
but much remains to be learned (Wilson and Conrad,
1984).
3.2 Physical Containment Techniques
3.2.7 Removal
The purpose of removing contaminated soil and
ground water, associated with a plume of
contamination, would be to treat and/or relocate the
wastes to a better engineered and controlled, or
environmentally more favorable disposal site.
Conceptually, removal and reburial of the
contaminated material to a more controlled situation
appears to solve the contamination problem. In
practice, however, there are many considerations to
deal with before excavation and reburial are used as a
remedial action technique. Considerations include (1)
excavation of bulky, partially decomposed or
hazardous wastes; (2) distance to acceptable reburial
site; (3) condition of roads between sites; (4)
accessibility of both sites; (5) political, social, and
economic factors associated with locating a new site;
(6) disposition of contaminated ground water; (7)
control of nuisances and vectors during excavation;
(8) reclamation of the excavated site; and (9) costs
(Tolman ef a/., 1978). Due to these considerations
and especially the cost of excavation, transportation
and new site preparations, removal and reburial
should be considered as a last resort or in cases of
severe pollution where cost is not significant
compared to the importance of the resource being
protected. In some cases removal and reburial in an
approved facility is simply transferring a problem from
one location to another.
3.2.2 Barriers to Ground-Water Flow
Subsurface barriers are designed to either prevent or
control ground water flow into or through desired
locations. The types of barriers used include slurry
trench walls, grout curtains, vibrating beam walls,
sheet piling, bottom sealing, block displacement and
passive interceptor systems (Knox ef a/., 1984;
Ehrenfeld and Bass, 1984).
A slurry trench wall is constructed by excavating a
vertical trench to a desired depth while throughout the
excavation process the trench is kept filled with a clay
slurry composed of a 5 to 7 percent by weight
suspension of bentonite in water. The bentonite slurry
maintains the vertical stability of the trench walls by
exerting a greater hydrostatic pressure against the
walls than the surrounding ground water, and also by
forcing bentonite into the pores of the soil in the
trench walls thus forming a low permeability layer of
soil and bentonite called a "filter cake."
As the slurry trench is being excavated it is
simultaneously being backfilled with an engineered
material that forms the final wall. The three major
types of slurry trench backfill mixtures are (1) soil
bentonite, (2) cement bentonite, and (3) concrete.
The type and ratios of backfill chosen depend upon
the specific site characteristics as well as the desired
properties of the slurry trench wall including
permeability, strength, compatibility with contaminants
and cost. Although costly, slurry trench walls are
generally the least expensive form of passive
ground-water barrier. A properly designed slurry
trench can lower the water table by providing a
complete seal down to a low permeability layer or by
increasing the length of the ground-water flow path
and thereby creating an energy loss.
Grouting can be defined as the pressure injection of a
stabilizing material into subsurface soils or rock in
order to fill, and thereby seal the voids, cracks,
fissures or other openings in the soil or rock strata.
Grout curtains are fixed, underground physical
barriers formed by injecting grout, either paniculate
(i.e., Portland cement) or chemical (i.e., sodium
silicate), though tubes which are driven into the
ground on two to three foot centers and withdrawn
slowly during injection. Two or more rows of grout are
normally needed to provide a good seal. Like a slurry
trench the grout curtains are normally emplaced down
to an impermeable layer. The rate of injection of the
grout material is determined by site-specific
characteristics. If the injection rate is too slow,
premature grout/soil consolidation occurs, and if the
rate is too fast, fracturing of the soil formation may
result. There are many available grouts; however, the
selection of grout material depends on site specific
factors such as soil permeability, soil grain size, rate
of ground water flow, chemical constituents of soil
and ground-water, required grout strength, and cost.
Because a grout curtain can be as much as three
36
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times as costly as a slurry wall, it is rarely used when
ground water has to be controlled in soil or loose
overburden. The major use of grout curtains is to seal
voids in porous or fractured rock where other
methods of ground water control are impractical.
A variation of a grout curtain is the vibrating beam
technique for emplacing thin (approximately 4 in)
curtains or walls. Although it is sometimes called a
slurry wall technique, it is more closely related to a
grout curtain since the slurry is injected through a
pipe similar to grouting. A suspended I-beam,
connected to a vibrating driver-extractor, is vibrated
through the ground to the desired depth. As the beam
is raised at a controlled rate, slurry is injected through
a set of nozzles at the base of the beam, filling the
void left by the withdrawal of the beam. The entire
process is repeated with subsequent placements of
the I-beam overlapping the previous placement to
provide continuity. The vibrating beam technique is
most efficient in loose, unconsolidated deposits such
as sands and gravels. Where suitable conditions exist
the vibrating beam technique has been used to
depths of 80 feet. Costs using the vibrating beam
technique are comparable to conventional slurry
trenching methods.
Sheet piling cutoff walls can be made of wood,
reinforced concrete or steel; however, steel sheet
piles represent the most effective material for
constructing a ground-water barrier. Construction of
a steel sheet pile cutoff wall involves driving lengths
of steel sheets through unconsolidated deposits with
a pile driver. The individual steel sheet piles are
connected along the edge of each pile through
various types of interlocking joints. These joints
provide permeable pathways for ground-water
movement which may or may not become watertight
naturally depending on the soil characteristics. It may
be necessary to fill these joints with an impermeable
material such as a grout; however, the ability to
ascertain the success ofx the grouting operation is
questionable. Steel is a readily corrodible material and
therefore the lifetime of the steel sheet piles is
dependent on the corrosive nature of the soil, ground
water, and contaminants with which the steel piles
come in contact. A common recommendation is that
steel pilings be chemically coated or electrically
protected so as to minimize corrosion. Although there
are limitations, sheet piling cutoff walls may be used
to contain contaminated ground water, divert a
contaminant plume to a treatment facility, and divert
ground-water flow around a contaminated area.
Block displacement is a new plume management
method where a slurry is injected in such a manner
that it forms a subsurface barrier around and below a
specific mass or "block" of earth. Continued pressure
injection of the slurry produces an uplift force on the
bottom of the "block" which results in a vertical
displacement proportional to the slurry volume
pumped, thus the name block displacement. This
technology is still in the developmental stages,
especially verification of the bottom barrier, so cost
data are not published.
Membrane and synthetic sheet curtains can be used
in applications similar to grout curtains and sheet
piling. The membrane is placed in a trench
surrounding or upgradient from the plume of interest,
thereby enclosing the contamination or diverting the
ground-water flow. Placing a membrane liner in a
slurry trench application has also been tried on a
limited basis. Attaching the membrane to an
impervious layer and having perfect seals between
sheets is difficult but necessary in order for
membranes and other synthetic sheet curtains to be
effective. Impermeable synthetic membranes have
also been used on the downgradient side of
interceptor trenches to stop the migration of
petroleum products for subsequent recovery.
Passive interceptor systems consist of trenches
excavated to a depth below the water table with the
possible placement of a collection pipe in the bottom
of the trench. These interceptor systems can be used
as preventive measures (i.e., leachate collection
systems), abatement measures (i.e., interceptor
drains), or in product recovery from a ground water
(i.e., oil, gasoline). Interceptor drains are generally
used to either lower the water table beneath a
contamination source or to collect ground water from
an upgradient source in order to prevent leachate
from reaching uncontaminated wells or surface water.
Interceptor systems are relatively inexpensive to
install and operate and provide a means for leachate
collection without impermeable liners. On the other
hand, interceptor systems are not well suited to
poorly permeable soils and the systems require
continuous and careful monitoring to assure adequate
leachate collection.
3.2.3 Surface Water Controls
Surface water control measures are used to minimize
the infiltration of surface water or direct precipitation
onto a waste site, thereby minimizing the amount of
leachate produced. There are three basic
technologies used to control surface water in a
particular area or waste site. The first is changing the
contour and runoff or runon characteristics of the site.
The second is providing a cover barrier to infiltration
by reducing the permeability of the land surface
(surface sealing or capping). The third is revegetating
the site so that the waste site cover material is
stabilized, seasonal evapotranspiration is increased,
and infiltration is decreased due to vegetation
interception of direct precipitation.
Changing the contour and runoff or runon
characteristics of a particular site can be
accomplished by several standard engineering
techniques. Some of the more common techniques
37
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include dikes and berms, ditches, diversions,
waterways, terraces, benches, chutes, downpipes,
levees, seepage basins, sedimentation basins, and
surface grading.
Surface sealing is accomplished by covering or
capping a waste site with a low permeable material to
prevent water from entering the site, thus reducing
leachate generation and also controlling vapor or gas
produced. Covers or caps can be constructed from
native soils, clays, synthetic membranes, soil cement,
bituminous concrete, certain waste materials, or
asphalt/tar materials. Capping is normally an
economical technique, and because the surface is
accessible, the cap can be monitored, maintained,
and repaired.
Revegetation may be a cost effective method to
stabilize the surface of a waste site, especially when
preceded by capping and contouring. Vegetation
reduces raindrop impact, reduces runoff velocity, and
strengthens the soil mass, thereby reducing erosion
by wind and water, and improves the site
aesthetically.
3.2.4 Limitations of Physical Containment
Along with the positive attributes, each of the physical
containment techniques have certain limitations.
Removal is an extremely expensive and difficult
procedure often plagued with political, social and
economic constraints. Construction of barriers to
ground-water movement can also present many
problems both site and technique related. Slurry walls
are limited by the availability of bentonite and the
patents associated with several aspects of the
construction procedures. Chemical grouts are
expensive, some grouting techniques are proprietary,
and grouting in general is limited to soils with
permeabilities 10"5 cm/sec or greater. Sheet piling is
not initially watertight, ineffective where large rocks
are present, and is subject to corrosion depending
upon site characteristics. Block displacement is an
untried technique in its infancy and needs more
verification studies especially concerning the bottom
barrier. Passive interceptor systems are not well
suited to slightly permeable soils and require
continuous monitoring and maintenance. Limitations
associated with surface water controls include
availability of cover material to develop contours,
availability of natural clay deposits for caps, expense
of manmade cover materials (concrete) and synthetic
membrane liners for caps, and initial time period and
cover required for vegetation.
The various limitations illuminate the fact that each
type of physical containment must be considered on
a case by case basis taking into consideration all the
many different site specific variables. It is possible
that even though a specific technique may be
expensive or the raw material may not be readily
available, when all site variables are considered, that
particular technique for physical containment may be
the only viable alternative.
3.3 Hydrodynamic Controls
Hydrodynamic controls are employed to isolate a
plume of contamination from the normal ground-
water flow regime in order to prevent the plume from
moving into a well field, another aquifer, or surface
water. Isolation of the contaminated plume is
accomplished when uncontaminated ground water is
circulated around the plume in the opposite direction
of the natural ground-water flow. The circulated
zone creates a ground-water (hydrodynamic) barrier
around the plume. Ground water upgradient of the
plume will flow around the circulated zone while
ground water downgradient will be essentially
unaffected.
3.3.1 Well Systems
Well systems are used for hydrodynamic control of
contaminated plumes by manipulating the hydraulic
gradient of ground water through injection and/or
withdrawal of water. The three general classes of well
systems include (1) well point systems, (2) deep well
systems, and (3) pressure ridge systems. All three
types of well systems may require the installation of
several wells at selected sites.
Well point systems consist of several closely-
spaced, shallow wells connected to a main header
pipe which is connected to a suction lift pump. Well
point systems are used only for shallow aquifers
because of the drawdown limitations as determined
by the static water level and the limits of the pump.
These systems should be designed so that the
drawdown of the system completely intercepts the
plume of contamination.
Deep wells are similar to well point systems except
they are used for greater depths and are normally
pumped individually. These wells are used in
consolidated formations where the water table is too
deep for economical use of suction life systems.
Since the maximum depth for suction lift is around 25
feet, deep wells normally employ jet ejector or
submersible pumps, or eductor well points.
Pressure ridge systems are produced by injecting
uncontaminated water into the subsurface, through a
line of injection wells, either up-gradient or down-
gradient from a plume of contamination. Up-gradient
ridges or mounds are used to force up-gradient
uncontaminated ground water to flow around a
contaminant plume while the contaminants are being
collected by a line of down-gradient pumping wells.
The procedure increases the velocity of ground water
into the plume and to the recovery wells, and serves
to wash the aquifer. Pressure ridge systems located
down-gradient are normally used in combination with
up-gradient pumping wells which supply
uncontaminated injection water. In either case the
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injection of fresh water produces an uplift or mound in
the original water table which acts as a barrier by
forming a ridge which pushes the contaminated
plume away from the mound.
Due to the economic incentives and the large number
of cases of hydrocarbon leakage from storage tanks,
considerable work has been directed toward applying
well system technology to hydrocarbon recovery.
Many variations of hydrocarbon recovery systems
have been proposed centered around single or
multiple pump systems and recovery wells. The type
of recovery well used is dependent upon site specific
characteristics and cost.
Hydrodynamic control systems offer a high degree of
design flexibility and compared to passive
containment can be easily constructed at minimal
expense. These are also moderate to high operational
flexibility which allows the system to meet increased
or decreased pumping demands. When rapid
response to a contamination problem is needed,
pumping and injection wells can be installed relatively
quickly as compared to certain passive barriers (i.e,
slurry trenches, grout curtains, etc.). If the
contamination threat is considered an emergency
condition, then a hydrodynamic control system may
be the temporary answer.
3.3.2 Limitations of Hydrodynamic Control
Even with the advantages discussed, well systems
are not a permanent panacea to a ground-water
contamination problem. Well systems simply are
methods to stop the migration of a plume until a
permanent solution can be decided upon. Some of
the more specific limitations include (1) higher
operation and maintenance costs than passive
barriers including electrical power and manpower, (2)
system failures, due to breakdown of equipment or
power outages, can lead to contaminant movement,
(3) flexibility is reduced in fine silty soils, and (4)
incorrect pumping rates can draw a significant portion
of the plume into the wells making treatment
necessary before recharge into the aquifer.
3.4 Withdrawal and Treatment
Withdrawal and treatment of contaminated ground
water is one of the most often used processes or
current technology for cleaning up aquifers. The type
of contamination and the cost associated with
treatment will determine what specific treatment
technology will be used. There are three broad areas
of treatment possibilities, namely (1) physical, which
includes adsorption, density separation, filtration,
reverse osmosis, air and stream stripping, and
incineration, (2) chemical, which includes
precipitation, oxidation/reduction, ion exchange,
neutralization, and wet air oxidation, and (3)
biological, which includes activated sludge, aerated
surface impoundments, land treatment, anaerobic
digestion, trickling filters, and rotating biological discs.
3.4.1 Physical
The two major adsorption methods receiving the
greatest attention as treatment methods are granular
(to include powdered activated carbon) activated
carbon (GAC) and synthetic resin adsorption. Both
GAC and resins remove dissolved contaminants from
water by adsorbing specific molecules. GAC is by far
the most widely used adsorbent because synthetic
resins are extremely costly and are still somewhat in
the developmental stages of normal use. Synthetic
resins trap contaminants within the chemical structure
of the resin whereas GAC traps contaminants within
the physical pore structure of the carbon. Typical
adsorption sytems, whether GAC or resin, consist of
a large vessel partially filled with adsorbent. There is
an inlet for contaminated water and an outlet for
treated water. Influent water enters and is in contact
with the adsorbent for a specified period of time and
then exits for collection, recharge or further treatment.
Often systems are arranged with several tanks in
parallel or in series to allow for the most efficient
treatment possible. Once the micropore surfaces of
the GAC are saturated with contaminants the GAC
must either be replaced or thermally regenerated.
GAC is an effective and reliable means of removing
low solubility organics and some metals and inorganic
species. It can be used for treating a wide range of
contaminants over a broad concentration range.
A part of many treatment operations for contaminated
ground water is the technique of density separation
where suspended solids and water are separated
depending upon their individual densities. If
suspended solids are present, often common
wastewater treatment operations such as clarifiers,
settling chambers, and sedimentation basins are
employed. Gravity separation is used when two-
phased aqueous wastes are present. Gravity
separation is a purely physical phenomenon in which
one phase (i.e., oil, hydrocarbons) is allowed to
separate from the other phase (i.e., water) in a
conical tank and then discharged accordingly.
Filtration is a physical process whereby suspended
solids are removed from solution by forcing the fluid
through a porous medium. The filter media consists
of a bed of granular particles (typically sand or sand
with anthracite or coal). Filters are often preceded by
sedimentation basins, and often precede biological or
activated carbon units in order to decrease the
suspended solids load. Filtration is a reliable and
effective means of removing low levels of solids
provided the solids content does not vary greatly and
the filter is backwashed at appropriate intervals.
In the process of osmosis a solvent spontaneously
flows from a dilute solution, through a semipermeable
membrane, to a more concentrated solution by
osmotic pressure. If enough pressure is placed on the
concentrated solution to overcome osmotic pressure
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then water will flow toward the dilute phase thereby
creating reverse osmosis. In reverse osmosis the
contaminants are allowed to build up in a circulating
bath on one side of the membrane while relatively
pure water passes through the membrane.
Basically a reverse osmosis unit is composed of the
membrane, a membrane support structure, a
containing vessel, and a high pressure pump with the
most critical elements being the membrane and the
membrane support structure. The process is used to
separate ions and small molecules in true solution
from water, and to decrease the dissolved solids
concentration, both organic and inorganic. Advances
in membrane technology have made it possible to
remove low molecular weight organics such as
alcohols, ketones, amines, and aldehydes. If
pretreatment measures are performed such as
removal of suspended solids, pH adjustments,
removal of oxidizers, oil, and grease, then reverse
osmosis has been shown to be an effective treatment
technology.
Stripping is a mass transfer process whereby volatile
contaminants are removed from aqueous wastes by
passing air or stream through the wastes. Air stripping
has been directly applied to ground- water treatment
in removing trichloroethylene (TCE), trihalomethane
(THM), and hydrogen sulfide. Removal rates as high
as 99 + percent for TCE from ground water, and 90 +
percent for ammonia from wastewater has been
observed.
Air Stripping is frequently accomplished in a stripping
lagoon or more commonly in a packed tower
equipped with an air blower. The packed tower works
on the principle of countercurrent flow. The water
stream flows down through the packing while the air
flows upward, and is exhausted through the top to the
atmosphere or to emission control devices (e.g.,
condensers, carbon adsorption filters). The volatile
substances tend to leave the aqueous stream for the
gas phase. In the cross-flow tower, water flows
down through the packing as in the countercurrent
packed column, however, the air is pulled across the
water flow path by a fan. The coke tray aerator
requires no blower. The water being treated trickles
through several layers of trays. This produces a large
surface area for gas transfer. Diffused aeration and
induced draft stripping use aeration basins or lagoons
similar to standard wastewater treatment technology.
Water flows through the basin from top to bottom or
from side to side with the air dispersed through
diffusers at the bottom of the basin. The air-to-
water ratio is significantly lower in the basins than in
either the packed column or cross-flow towers.
Temperature has an effect on the mass transfer
coefficient of substances. This is an important point
when contaminated ground water contains
compounds that are very soluble (i.e., compounds
with low Henry's Law constants). High water solubility
makes their removal by ambient temperature air
stripping almost impossible. It has been shown that
removal efficiency increases dramatically with
temperature and less sharply with the air-to-water
ratio. Therefore, high temperature air stripping, or
steam stripping, offers increased flexibility and should
be investigated for each case as necessary.
Incineration is a treatment method which employs
high temperature oxidation under controlled conditions
to decompose a substance into products that
generally include COa, H20 vapor, S02, NOX, HCI,
and products of incomplete combustion require air
pollution control equipment to prevent release of
undesirable species into the atmosphere. Incineration
methods can be used to destroy organic
contaminants in liquid, gaseous and solid waste
streams.
The most common incineration technologies are liquid
injection, rotary kiln, fluidized-bed, and multiple
hearth. Rotary kiln and multiple hearth incinerators
can be used with most organic wastes including
solids, sludges, liquids and gases, while liquid
injection incinerators are limited to pumpable slurries
and liquids. Fluidized-bed incinerators work well for
organic liquids, gases and granular or well processed
solids. Incineration offers one of the most effective
technological methods for complete destruction of
organic compounds.
3.4.2 Chemical
Contaminated ground water can be withdrawn and
treated chemically by various techniques. Among the
more common chemical treatment technologies are
neutralization, precipitation, oxidation and reduction,
ion exchange, and chemical fixation.
Neutralization is merely a process whereby an acid or
base is added to a waste in order to adjust the pH.
Neutralization is a relatively simple unit process which
can be performed using ordinary and commonly
available treatment equipment. It is often used prior to
other treatment processes where the pH of the waste
is critical (e.g., biological treatment and carbon
adsorption).
Precipitation is a physiochemical process whereby a
substance in solution is transformed to the solid
phase. Precipitation can be accomplished by (1)
adding a chemical that will react with the contaminant
in a solution forming a sparingly soluble compound,
(2) adding a chemical which changes the solubility
equilibrium of a waste thus reducing the solubility of
the specific contaminants, and (3) changing the
temperature to decrease the solubility of the
contaminants. Removal of metals as carbonates,
hydroxides, or sulfides is the most common
application of precipitation in wastewater treatment.
Many precipitation reactions (e.g., metal sulfides) do
not readily form floe (large fluffy precipitates)
particles, but rather precipitates very fine and
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relatively stable colloidal particles. In these cases
flocculating agents (e.g., alum and/or polyelectrolytes)
must be added to cause flocculation of the metal
sulfide precipitates. The effectiveness of
precipitation/flocculation reactions is dependent upon
the nature and concentration of the contaminants,
and upon the process design. The process design
must consider the optimum chemicals and dosages,
suitable chemical addition systems, optimum pH and
mixing requirements, sludge production, and sludge
flocculation, settling and dewatering characteristics.
Oxidation/reduction processes are employed to raise
(oxidation) or lower (reduction) the oxidation state of a
substance or substances in order to reduce toxicity or
solubility, or to transform the substance to a form
which can be more easily handled. Commonly used
reducing agents include sulfite salts, sulfur dioxide
and the base metals (i.e., iron, aluminum and zinc).
Chemical reduction is used primarily for the reduction
of hexavalent chromium, mercury and lead. There are
currently no practical applications involving reduction
of organic compounds. Oxidation, however, has found
extensive use in treatment of organic wastes.
Oxidizers which are most often used in wastewater
treatment include oxygen or air, ozone, ozone with
ultraviolet light, chlorine gas, hypochlorites, chlorine
dioxide, and hydrogen peroxide.
Ion exchange is a process whereby toxic ions are
removed from the aqueous phase by being
exchanged with relatively harmless ions held by the
ion exchange material. Ion exchange is used to
remove a broad range of ionic species from water to
include (1) all metallic elements when present as
soluble species, either anionic or cationic, (2)
ionorganic ions such as halides, sulfates, nitrates,
cyanides, etc., (3) organic acids such as carboxylics,
sulfonics, and some phenols, and (4) organic amines
in sufficient acidity to form the acid salt (De Renzo,
1978). Ion exchange systems will function well in
dilute waste streams of variable composition provided
the effluent is monitored to determine when ion
exchange resin bed exhaustion has occurred.
Solidification/stabilization technologies reduce
leachate production potential by physically and/or
chemically binding a waste in a solid matrix. Wastes
are mixed with a binding agent to produce a solid
form. Solidification/stabilization processes include (1)
cementation, using Portland cement, (2) pozzolanic
cementation, (3) thermoplastic binding, (4) organic
polymer binding, (5) surface encapsulation, and (6)
glassification. Cementation and pozzolanic
cementation are generally the most widely applicable
to a wide range of waste compositions. Most
solidification/stabilization technologies are designed
for inorganic wastes and can be seriously affected by
high concentrations of organic wastes (i.e., increased
cure, set inhibition, flashset, etc.). However, research
is currently being conducted into specific interference
effects caused by particular types of wastes, and into
the solidification/stabilization of certain organic
wastes. Important waste characteristics that impact
solidification/stabilization processes include pH, buffer
capacity, water content, organic concentrations, and
specific inorganic constituents.
3.4.3 Biological
The function of biological treatment is to remove
organic matter from the waste stream through
microbial degradation. A number of biological
treatment processes exist which may be applicable to
the treatment of contaminated ground water, including
various forms of activated sludge, surface
impoundments, trickling filters, rotating biological
discs, fluidized bed reactors, land treatment, and
anaerobic digestion.
Activated sludge treatment consists of passing the
contaminated waste stream through an aeration basin
where it is aerated for several hours. During this time
an active microbial population develops which
degrades organic matter in the waste. In the process
a portion of the activated sludge is recycled and along
with new developing cells the microbial population is
maintained for further degradation of the waste
stream. Various versions of the activated sludge
process (e.g., oxygen, oxygen-enriched, extended
aeration, contact stabilization, etc.) are simply a result
of the type of aeration, time of aeration, and the
contact time with the activated sludge.
Surface impoundments or lagoons are similar to
activated sludge units without sludge recycle. Surface
impoundments are similar to a natural eutrophic lake
in that natural processes of microbial oxidation,
photosynthesis, and sometimes anaerobic digestion
combine to degrade organic wastes. Aeration may be
supplied passively by wind action or, in aerated
surface impoundments, by mechanical aerators.
Trickling filters are a form of biological treatment in
which a liquid waste (<1% suspended solids) is
trickled over a bed of rocks or synthetic material upon
which a slime layer of microbial organisms develops.
The microbes in the slime layer metabolize the
organics in the waste while oxygen to the
microorganisms is provided as air moves
countercurrent to the water flow.
A modification of the trickling filter is the biological
tower. The medium (e.g., of polyvinyl chloride,
polyethylene, polystyrene, or redwood) is stacked into
towers which typically reach 16 to 20 feet. The
contaminated water is sprayed across the top of the
tower and, as it moves downward, air is pulled
upward through the tower. A slime layer of
microorganisms develops on the media and removes
the organic contaminants as the water flows over the
slime layer.
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Another fixed film biological treatment process, similar
in operating principle to trickling filters, is the rotating
biological disc system. This system consists of a
series of rotating discs, connected by a shaft, set in a
basin or trough. Approximately 40 percent of each
disc's surface area is submerged in the basin and as
the contaminated water passes through the basin, the
microorganisms growing on the disc metabolize the
organics in the wastewater. As the discs rotate the
microorganisms are brought in contact with the air
where oxygen is obtained for growth.
Yet another fixed film process is the fluidized bed
reactor. Particles of substances such as sand or coal
are fluidized by the action of the aeration gas st ream
and the wastewater stream. These particles support a
dense growth of microorganisms which give rapid
treatment to the wastewater. This process is still
largely experimental.
Land treatment is the mixing or dispersion of wastes
into the upper zone of the soil-plant system with the
objective of microbial stabilization, adsorption, and
immobilization leading to an environmentally
acceptable degradation of the waste. The four major
land treatment options are: (1) irrigation, (2) overland
flow, (3) infiltration-percolation, and (4) leachate
recycle. Before waste is applied, the assimilative
capacity of the land treatment plant-soil system
must be determined for each contaminant present,
considering the nature of the pollutant (i.e.,
biodegradability, mobility, uptake, toxicity, etc.) and
also the site characteristics (i.e., soil type,
hydrogeology, meteorology).
Anaerobic digestion is another biological treatment
process which can be used for organic contaminant
degradation. Whether anaerobic lagoons or anaerobic
digesters (totally enclosed) are used, the anaerobic
process merits consideration due to ease of
operation, minimal sludge production, and energy
efficiencies.
3.4.4 Limitations ol Withdrawal and Treatment
Techniques
As with any treatment process, the importance of
limitations associated with each process is
determined by the urgency of treatment, the
importance of the resource, and the availability of
funding for the treatment. Some of the more
important limitations characteristic of the previously
mentioned physical, chemical, and biological
treatment processes include:
Physical
1. Carbon adsorption is intolerant of high suspended
solids; can be poisoned by high heavy metals
concentrations; requires pretreatment for oil and
grease >10 ppm; and has high operation and
maintenance costs.
2. Resin adsorption is more expensive and usually
has less capacity than carbon adsorption; resin is
intolerant of strong oxidizing agents and
suspended solids.
3. Density separation often yields incomplete
removal of hazardous compounds and generates
large quantities of contaminated sludges.
4. Filter clogging (in filtration process) due to high
suspended solids causes reduced run lengths
and requires frequent backwashing or
replacement of the filter.
5. Reverse osmosis units are subject to chemical
attack, fouling, and plugging, and can require
extensive pretreatment.
6. Stripping is sensitive to pH, temperature, and
fluxes in hydraulic load; may be cost prohibitive at
temperatures below freezing; and may cause air
pollution problems.
7. Incineration may require thickening and
dewatering pretreatment; may pose air pollution
problems; produces an inorganic ash (possibly
hazardous); and may require costly fuel or power
for operation.
Chemical
1. Precipitation can be limited by the presence of
complexing agents in the waste and the
precipitate itself may be a hazardous waste.
2. Reduction is used primarily for reducing
hexavalent chromium, mercury and lead. There
are no current applications for reducing organic
compounds.
3. Chemical oxidation costs are generally higher
than biological treatment. Some organics are
resistant to most oxidants and in some cases
partial oxidation generates toxic compounds.
4. The effectiveness of ion exchange is reduced by
high suspended solids and/or high concentrations
of certain organics.
5. Design considerations should be made to
accommodate the corrosive nature of some
neutralization reagents.
6. Solidification/stabilization techniques result in
increased volume and weight for disposal and are
still subject to leaching of contaminants. Certain
wastes cause interferences with the
solidification/stabilization processes.
Biological
1. Activated sludge costs are high with intensive
operation and maintenance costs; is sensitive to
suspended solids and metals; generates sludge
high in metals and refractory organics; and is
fairly energy intensive.
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2. Surface impoundments are sensitive to shock
loadings and temperature effects. Gas generation
and chemical volatilization are problems with
anaerobic lagoons.
3. Trickling filters, rotating biological discs, and
fluidized bed reactors require an energy source
and are vulnerable to below freezing
temperatures; have potential for odor problems;
require long recovery times if disrupted; and have
limited flexibility (i.e., minimize variations in
operating conditions such as flow and
composition).
4. Land treatment requires large areas of land; has
the potential for ground-water contamination;
and the ground water must be monitored.
3.5 In-Situ Treatment Techniques
3.5.1 Chemical/Physical
In-situ physical and chemical techniques, for the
most part, are not well developed and are highly
dependent on a number of physical factors, including
aquifer permeability and the nature of specific
contaminants, as well as the natural geochemistry of
the earth materials.
In-situ techniques are certainly more aesthetically
desirable than most other alternatives since they
require minimal surface facilities and minimize public
exposure to pollutants. Costs are quite variable and
directly relate to the contaminant constituents present
their required control agents, hydrogeologic
conditions of the aquifer, aerial extent of the pollution
source and physical accessibility to the site.
In-situ chemical detoxification techniques include
injecting neutralizing agents for acid or caustic
leachates, adding oxidizing agents to decompose
organics or precipitate inorganic compounds, adding
agents that promote other natural degradation
processes, bonding contaminants, and immobilization
or reaction in treatment beds. These techniques
should only be considered in cases where specific
contaminants, their concentration levels and the
extent of the contamination plume in the aquifer are
well defined. The treatment agents are specific for the
class of contaminant. For example, metals may be
rendered insoluble and immobile with alkalines or
sulfides, and cyanides can be oxidized using strong
oxidizing agents such as sodium hypochlorite or by
encapsulation in an insoluble matrix. Cations may be
precipitated by injecting various anions or by in-situ
aeration. Hexavalent chromium could be made
insoluble by injecting specific reducing agents. Free
fluorides can be insolubilized by the injection of
solutions containing the calcium ion.
In-situ physical/chemical treatment processes
generally entail installing a series of wells for
chemical injection at the head of or within the plume
of contaminated ground water. An alternate technique
that has been used in shallow aquifers is in-situ
permeable treatment beds. These are often used to
detoxify migrating leachate plumes in ground water
from landfills. Trenches are filled with a reactive
permeable medium; contaminated ground water
entering the trench reacts with the medium to
produce a nonhazardous soluble product or a solid
precipitate. Among the materials commonly used in
permeable bed trenches are limestone to neutralize
acidic ground water and remove heavy metals;
activated carbon to remove nonpolar contaminants
such as carbon tetrachloride (CCU), polychlorinated
biphenyls (PCBs) and benzene by adsorption; and
zeolites and synthetic ion exchange resins for
removing solubilized heavy metals.
Permeable treatment beds are applicable only in
relatively shallow aquifers because the trench must
be constructed down to the level of the bedrock or an
impermeable clay. They are also often effective for
only a short time because they lose their reactive
capacity or become plugged with solids. Over-
design of the system or replacement of the
permeable medium can lengthen the time period over
which permeable treatment is effective.
There are a number of disadvantages associated with
both of these techniques. In permeable treatment
beds, plugging of the bed may divert contaminated
ground water, or channeling through the bed may
occur. Changing hydraulic loads and contaminant
levels may mean that detection times in the beds are
inadequate. In the chemical injection technique,
pollutants may be displaced to adjacent areas when
chemical solutions are injected under higher hydraulic
heads. In addition, hazardous compounds may be
produced by reaction of injected chemical solutions
with waste constituents other than the treatment
target.
Mobilization of contaminants by injecting surfactants
during soil washing is possible. Surfactants and
alkaline flooding for enhanced secondary oil recovery
are being used experimentally with moderate
success. Most oil field surfactants are expensive
refined biodegradable organics, while alkaline floods
produce lye. This approach does not appear
promising for aquifer restoration because of the
addition of potentially hazardous materials or the
creation of hazardous degradation by-products
which would then have to be dealt with.
To recover hydrocarbons, there are three possible
physical-chemical methods. At shallow depths,
thermal or steam floods may be helpful. On a larger
scale, alcohol flooding may be feasible. Alcohol is
easily produced and dissolves the hydroca bon.
Theoretically, if an entire polluted zone is flooded with
alcohol, all of the residual hydrocarbon can be
removed. Limitations of with this method include high
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cost, phase-behavior difficulties and lack of field
experience.
Other in-situ treatment techniques have been
suggested, including radio frequency in-situ heating
or in-situ vitrification, using an electric current to
melt the soils and waste in place. The economics for
field application of these systems are unknown.
3.5.2 Biodegradation
3.5.2.1 Natural Biological Activity in the
Subsurface
In-situ biorestoration of the subsurface is a relatively
new technology that has recently gained considerable
attention. Scarcely more than a decade ago,
conventional wisdom assumed that the subsurface
below the zone of plant roots was, for all practical
purposes, sterile.
Recent research has indicated that the deeper
subsurface is not sterile, but in fact, harbors
significant populations of microorganisms. Bacterial
densities around 106 organisms/g dry soil have been
found in several noncontaminated aquifers. The
water-table aquifers examined so far exhibit
considerable variation in the rate of biodegradation of
specific contaminants that enter the subsurface
environment. Rates can vary two to three orders of
magnitude between aquifers or over a vertical
separation of only a few feet in the same aquifer.
Although extremely variable, the rates of
biodegradation are fast enough to protect ground -
water quality in many aquifers.
Although they are not clearly defined, several
environmental factors are known to influence the
capacity of indigenous microbial populations to
degrade contaminants. These factors include
dissolved oxygen, pH, temperature, oxidation-
reduction potential, availability of mineral nutrients,
salinity, soil moisture, the concentration of specific
pollutants, and the nutritional quality of dissolved
organic carbon in the ground water.
Many water-table aquifers contain oxygen, which
can support aerobic microorganisms that can degrade
a wide variety of organic contaminants. Examples
include acetone, isopropanol, methanol, ethanol, t-
butanol, benzene, toluene, the xylenes, and other
alkybenzenes that leak into ground water from
gasoline spills or solvent spills (Novak, et a/., 1984;
Lokke, 1984; Jhaveri and Mazzacca, 1983; Wilson ef
a/., 1986; Lee ef a/., 1984); napthalene, the
methylnaphthalenes, fluorene, acenaphthene,
dibenzofuran and a variety of other polynuclear
aromatic hydrocarbons released from spilled diesel oil
or heating oil (Wilson, et a/., 1985); and many
methylated phenols and heterocyclic organic
compounds seen in certain industrial waste waters.
Many synthetic organic compounds can also be
degraded. Examples include dichlorobenzenes (Kuhn,
et at., 1985), the mono-, di-, and trichlorophenols
(Suflita and Miller, 1985), the detergent builder
nitrilotriacetic acid (NTA) (Ward, 1985), and some of
the simpler chlorinated compounds such as
methylene chloride (dichloromethane) (Jhaveri and
Mazzacca, 1983).
The extent of biodegradation of these compounds in
ground water is limited by the concentration of
oxygen. For the compounds discussed above, roughly
two parts of oxygen are required to completely
metabolize one part of organic compound. For
example, microorganisms in a well-oxygenated
ground water containing 4 mg/l of molecular oxygen
can degrade only 2 mg/l of benzene. The solubility of
benzene (1780 mg/l) is much greater than the
capacity of its aerobic degradation in ground water.
Obviously, the prospects for aerobic metabolism of
these compounds will depend on their concentration
as well as on the concentration of other degradable
organic materials in the aquifer. Concentrated plumes
of organic contaminants cannot be degraded
aerobically until dispersion or other processes dilute
the plume with oxygenated water.
Many of the commonly encountered organic
pollutants in aquifers are synthetic organic solvents
that are very persistent in oxygenated waters.
Examples include tetrachloroethylene (PCE),
trichloroethylene (TCE), cis and trans 1,2-
dichlroethylene, ethylene dichloride (1,2-
dichloroethane), 1,1,1-trichloroethane (TCA), 1,1,2-
trichloroethane, carbon tetrachloride, and chloroform.
This important class of organic contaminants
commonly enters ground water as spills from
underground storage tanks. Ground-water
contamination in the Santa Clara Valley of California
(Silicon Valley) is a good example. Recent research
has shown that this class of organic contaminants
can be cometabolized by bacteria that grow on
gaseous aliphatic hydrocarbons like methane or
propane. The potential use of cometabolism for in situ
restoration is under evaluation.
When the concentration of organic contaminants is
high, oxygen in the ground water will be totally
depleted and aerobic metabolism will stop. However,
further biotransformations often will be mediated by a
variety of anaerobic bacteria. Anaerobes that produce
methane, called methanogens, are only active in
highly reduced environments. Molecular oxygen is
very toxic to them. Methane can be produced by the
fermentation of a few simple organic compounds
such as acetate, formate, methanol, or methylamines.
Molecular hydrogen can also be used in the reduction
of inorganic carbonate to methane. Although the
microorganisms that actually produce the methane
can use a very limited set of organic compounds,
they can act in consort with other microorganisms
which break more complex organic compounds down
to substances that the methanogenic organisms can
44
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use. These partnerships or consortia can totally
degrade a surprising variety of natural and synthetic
organic compounds.
The rates of reaction are usually slow and often
require long lag periods before active transformation
begins (Wilson, 1985). Microbiologists are
accustomed to microorganisms that grow to high
densities in only a few days, and rarely conduct
experiments that last longer than a few weeks.
However, the residence time of organic pollutants in
aquifers is at least months or years and is frequently
decades to centuries. As a result, much of what was
learned in earlier laboratory studies cannot be applied
to the subsurface environment. Currently,
microbiologists are re-examining the potential for
biodegradation of organic contamination in ground
waters that actively produce methane and are finding
many expected reactions.
It was previously thought that the metabolism of
benzene, toluene, the xylenes and other
alkylbenzenes required molecular oxygen because
oxygen is a co-substrate for the only known enzyme
that can begin the metabolism of this class of
compounds (Young, 1984). Thus, their metabolism
would not be expected in methanogenic
environments. Yet recently, the metabolism of these
compounds was demonstrated in methanogenic river
alluvium that has been contaminated with landfill
leachate (Wilson and Rees, 1985). When radioactive
toluene was added to this material at least half the
carbon was metabolized completely to carbon dioxide.
The same materials also metabolized several methyl
and chlorophenols (Suflita and Miller, 1985). Very
recently extensive anaerobic metabolism of
alkylbenzenes has been demonstrated in a sandy
water-table aquifer contaminated with aviation
gasoline released from an undergound storage tank.
The halogenated solvents that are presistent in
oxygenated ground water can be transformed in
methanogenic ground water. Examples include
trichloroethylene, tetrachloroethylene, the
dichloroethylenes, 1,1,1-dichloroethane, carbon
tetrachloride and chloroform (Parsons, et a/., 1984;
Parsons, et a/., 1985; Wood et a/., 1985). Ethylene
dibromide is also transformed (Wilson and Rees,
1985). The chlorinated ethylenes undergo a
sequential reductive dehalogenation from
tetrachloroethylene to trichloroethylene, then to the
dichloroethylenes (primarily the CIS isomer) and
finally to vinyl chloride (Wood, et at., 1985). In some
subsurface environments, appreciable quantities of
vinyl chloride accumulate, which is unfortunate
because this compound is considerably more toxic
and carcinogenic than its parent compound. In other
subsurface environments the vinyl chloride is further
metabolized. The factors that control the fate of vinyl
chloride are unknown (Wilson, 1985). The
chloroalkanes follow a similar pattern (Wood, et a/.,
1985); carbon tetrachloride is converted to
chloroform, then to methylene chloride, while 1,1,1-
dichloroethane is converted to 1,1-dichloroethane,
which in turn goes to ethyl chloride.
These reductive dehalogenations resemble
respirations. In aerobic respiration, molecular oxygen
accepts an electron and is reduced to the
hydrogenated compound, water. The chlorinated
compounds accept electrons and are reduced to the
corresponding hydrogenated compound, while
chlorine is released as a chloride ion. The source of
electrons can be a co-occuring contaminant, such
as volatile fatty acids in landfill leachate, or it can be a
geological material. Reductive dechlorination of
trichloroethylene has been associated with flooded
surface soil, buried soils in glaciated areas, buried
layers of peat, and coal seams.
If oxygen is depleted, but conditions do not favor the
methanogens, certain classes of organic compounds
can be degraded by bacteria that respire nitrate or
sulfate. Ground waters recharged through soils that
support intensive agriculture often have high
concentrations of nitrate, and ground waters with
appreciable concentrations of sulfate are widespread,
particularly in arid regions. Microorganisms respiring
nitrate can degrade a number of phenols and cresols
(methylphenols). Recently, it has been shown that
nitrate respiring organisms in river alluvium could also
degrade all three xylenes (dimethylbenzenes) (Kuhn,
ef a/., 1985). Nitrate-respiring microorganisms can
also degrade carbon tetrachloride and a variety of
brominated methanes. However, they have not been
shown to degrade chloroform or those chlorinated
ethylenes or ethanes which are also stable in
oxygenated ground water (Bouwer and McCarty,
1983).
Like the methanogens, the sulfate-respiring bacteria
can participate in consortia that degrade a wide
variety of natural organic compounds. In contrast to
the behavior of methanogenic subsurface material,
chlorinated derivatives of naturally-occurring
aromatic compounds were not degraded in river
alluvium containing appreciable sulfate concentrations
(200 mg/l) and exhibiting active sulfate respiration
(Suflita and Miller, 1985, Suflita and Gibson, 1985).
As they did under highly reduced conditions,
tetrachloroethylene and trichloroethylene underwent
reductive dehalogenations.
As these studies have shown, natural biorestoration
does occur. Contaminants in solution in ground water
as well as vapors in the unsaturated zone can be
completely degraded or can be transformed to new
compounds. Undoubtedly, thousands of
contamination events are remediated naturally before
the contamination reaches a point of detection.
However, methods are needed to determine when
natural biorestoration is occurring, the stage the
restoration process is in, whether enhancement of the
45
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process is possible or desirable, and what will happen
if natural processes are allowed to run their course. A
number of researchers are presently working in this
area.
3.5.2.2 Enhanced Biorestoration - Basic
Principles
In subsurface situations, the populations of
metabolically-capable organisms increase until they
are limited by some metabolic requirement such as
food or mineral nutrients, or oxygen in the case of
aerobic organisms. Once this point is reached, the
rate of transformation of an organic material is
controlled by transport processes that supply the
limiting nutrient.
The vast majority of microbes in the subsurface are
firmly attached to soil particles. As a result, nutrients
must be brought by advection or diffusion through the
mobile phases, water and soil gas. In the simplest
and perhaps most common case, the organic
compound to be consumed for energy and cell
synthesis is brought in aqueous solution in infiltrating
water. At the same time oxygen is brought by
diffusion through the soil gas. In the unsaturated
zone, volatile organic compounds can also move
readily as vapors in the soil gas. Below the water
table all transport must be through liquid phases and
as a result the prospects for aerobic metabolism is
severely limited by the very low solubility of oxygen in
water. In the final analysis, the rate of biological
activity is controlled by:
o The stoichiometry of the metabolic process
o The concentration of the required nutrients in the
mobile phases
o The advective flow of the mobile phases or the
steepness of concentration gradients within the
phases
o Opportunity for colonization in the subsurface by
metabolically capable organisms
o Toxicity exhibited by the waste or a co-occurring
material.
3.5.2.3 Enhanced Biorestoration - Current
Practice
Most enhanced in-situ bioreclamation techniques
available today are variations of techniques pioneered
by Richard Raymond, Virginia Jameson, and co-
workers at Suntech during the period 1974-1978.
Suntech's process received a patent entitled
"Reclamation of Hydrocarbon Contaminated Ground
Waters" (Raymond, 1974). This process reduces
hydrocarbon contaminants in aquifers by enhancing
the indigenous hydrocarbon-utilizing microflora.
Nutrients and oxygen are introduced through injection
wells and circulated through the contaminated zone
by pumping one or more producing wells. The
increased supply of nutrients and oxygen stimulates
biodegradation of the hydrocarbons. Oxygen is
supplied by sparging air into the injection wells.
Raymond's process has been used largely to clean
up gasoline contaminated aquifers.
The first basic step in Raymond's process is usually
to employ physical methods to recover as much of
the gasoline product as possible. While the product is
being recovered, Raymond requires a detailed
investigation of the hydrogeology and the extent of
contamination. A laboratory study is conducted to
determine if the native microbial population can
degrade the contaminants. Laboratory studies also
identify the combination of minerals that promotes
maximum cell growth on the contaminant in 96 hours
under aerobic conditions at the ambient ground-
water temperature.
Considerable variation in the nutrient requirements
has been noted by Suntech. One aquifer required
only the addition of nitrogen and phosphorus, while
the growth of microbes in another aquifer was
stimulated best by the addition of ammonium sulfate,
mono- and disodium phosphate, magnesium sulfate,
sodium carbonates, calcium chloride, manganese
sulfate and ferrous sulfate. They found that chemical
analysis of the ground water was not helpful in
estimating the nutrient requirements of the system.
After the microbial laboratory investigations have
established the optimal conditions for growth of the
indigenous microbial population, the systems for
injecting the mixture of nutrients and oxygen and for
producing water to circulate them in the formation are
designed and built. Controlling the ground-water flow
is critical to moving oxygen and nutrients to the
contaminated zone and optimizing the degradation
process.
The Suntech process is reported to have met with
reasonable success when applied to gasoline spills in
the subsurface. Some of the sites treated by this
technique have been cleaned to the point where no
dissolved gasoline was present in the ground waters
and state regulatory standards were satisfied. The
State agencies in charge of cleaning-up other sites,
however, have directed operations to continue until
no trace of liquid gasoline can be detected. Most of
the sites have implemented appropriate ground-
water monitoring programs following clean-up. The
overall percent removal of total hydrocarbons using
this method has usually ranged from 70 to 80
percent.
The Suntech process does not provide for treatment
of the material above the water table. Soils or
geological material contaminated by leaking
underground storage tanks may be physically
removed during the process of removing the tank, in
which case the contaminated material can be
disposed of in an approved manner. However, in
cases where the extent of the pollution is large or the
46
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water table extends to a depth where physical
removal of contaminated material is totally impractical,
alternative methods are used. One of these methods
is construction of one or a series of surface infiltration
galleries. These galleries are used to recirculate
water, which has been treated, back through the
contaminated unsaturated zone. Oxygen is generally
added to the infiltrated water during an in-line
stripping process for volatile organic contaminants or
through aeration devices placed in the infiltration
galleries. Recirculation of the water also facilitates
movement of contaminant's to the recovery well. The
dislodged or solubilized contaminant can be treated in
a surface treatment system before the water is
reinjected.
In constructing an infiltration gallery, the most critical
factor is the rate of water infiltration. Silty and shaley
materials accept water very slowly. The site must be
tested and evaluated to determine size and
configuration of infiltration pits.
Whether the material is situated above or below the
water table, the rate of bioreclamation in hydrocarbon
contaminated zones is effectively the rate of supply of
oxygen. Table 3-1 compares the number of times
that water in contaminated material below the water
table, or air in material above it, must be replaced to
totally reclaim subsurface materials of various
textures. The calculations assume typical values for
the volume occupied by air, water and hydrocarbons
(De Pastrovich, et a/., 1979; Clapp and Hornberger,
1978). The actual values at a specific site will
probably be different. The calculations further assume
that the oxygen content of the water was 10 mg/l, that
of the air 200 mg/l, and that the hydrocarbons were
completely metabolized to carbon dioxide.
It is obvious that prodigious volumes of water are
needed in the finely-textured subsurface materials.
This has prompted a search for some mechanism to
increase the concentration of oxygen. The most
obvious approach is to sparge the injection wells with
Table 3-1 Estimated Volumes of Water or Air Required to Completely Renovate Subsurface Material that Contained
Hydrocarbons at Residual Saturation.
oxygen instead of air. This will increase the oxygen
concentration about five-fold. The water can also be
supplemented with hydrogen peroxide (Brown et a/.,
1984).
Laboratory studies have shown that hydrocarbon-
degrading bacteria can adapt to tolerate hydrogen
peroxide equivalent to 200 mg/l oxygen, a twenty-
fold increase in oxygen supply over water sparged
with air (Lee and Ward, 1985). However, the rate of
decomposition of hydrogen peroxide to oxygen must
be controlled. Rapid decomposition of only 100 mg/l
of peroxide will exceed the solubility of oxygen in
water resulting in bubble formation which could lead
to gas blockage and loss of permeability. Iron
catalyzes the decomposition of hydrogen peroxide in
ground water. Standard practice is to add enough
phosphate to the recirculated water to precipitate the
iron. Some suppliers add an organic catalyst that will
decompose the peroxide at a rate appropriate to the
rate of infiltration, so that the oxygen demand of the
bacteria attached to the solids is balanced by the
oxygen supplied by decomposing peroxide in the
recirculated water.
Obviously, successful use of hydrogen peroxide
requires careful control of the geochemistry and
hydrology of the site. In addition to the factors
mentioned already, hydrogen peroxide can mobilize
metals such as lead and antimony; and if the water is
hard, magnesium and calcium phosphates can
precipitate and plug up the injection well or infiltration
gallery.
3.6 Treatment Trains
In most contaminated hydrogeologic systems a
remediation process may be so complex in terms of
contaminant behavior and site characteristics that no
one system or unit is going to meet all requirements.
Very often, it is necessary to combine several unit
operations, in series and sometimes in parallel into
Proportion of Total Subsurface Volume
Occupied by:
Texture
Stone to Coarse Gravel
Gravel to Coarse Sand
Coarse to Medium Sand
Medium to Fine Sand
Fine Sand to Silt
Hydrocarbons
(when drained)
0.005
0.008
0.015
0.025
0.040
Air
(when drained)
0.4
0.3
0.2
0.2
0.2
Water
(when flooded)
0.4
0.4
0.4
0.4
0.5
Volumes Required
to meet
Hydrocarbons
Oxygen Demand
Air
250
530
1,500
2,500
4,000
Water
5,000
8,000
15,000
25,000
32,000
47
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one treatment process train in order to effectively
restore ground-water quality to a required level.
Barriers and hydrodynamic controls alone merely
serve as temporary plume control measures.
However, hydrodynamic processes must also be
integral parts of any withdrawal and treatment or in-
situ treatment measures.
Most remediation projects, where enhanced
biorestoration has been applied, have started by
removing heavily contaminated soils. This was usually
followed by installing pumping systems, to remove
free product floating on the ground water, before
biorestoration enhancement measures were initiated
to degrade the more diluted portions of the plume.
As noted earlier, there are numerous proven surface
treatment processes available for treating a variety of
organic and inorganic wastewaters. However,
regardless of the source of ground-water
contamination and the remediation measures
anticipated, the limiting factor is getting the
contaminated subsurface material to the treatment
unit or units, or in the case of in-situ processes,
getting the treatment process to the contaminated
material. The key to success is a thorough
understanding of the hydrogeologic and geochemical
c. ^racteristics of the area. Such an understanding
will permit full optimization of all possible remedial
actions, maximum predictability of remediation
effectiveness, minimum remediation costs, and more
reliable cost estimates.
3.7 Institutional Limitations on
Controlling Ground-Water Pollution
The principal criteria for selecting remediation
procedures should be the water quality level to which
an aquifer should be restored and the most economic
technology available to reach that quality level.
Unfortunately, there are numerous institutional
limitations that sometimes override these criteria in
determining if, when, what, and how remediation will
be selected and carried out.
Response to a ground-water contamination problem
is likely to require compliance with several local, state
and federal pollution control laws and regulations. If
the response involves handling hazardous wastes,
discharging substances into the air or surface waters,
or the underground injection of wastes, federal
pollution laws apply. These laws do not exempt the
activities of federal, state, or local officials or other
parties attempting to remediate contamination events.
They apply to generators and responding parties
alike, and it is not unusual for these pollution control
laws to conflict. For example, a hazardous waste
remediation project may be slowed, altered or
abandoned by the imposition, upon the party
undertaking the effort, of elaborate RCRA permit
requirements governing the transport and disposal of
hazardous wastes.
In-situ remediation procedures may be subject to
permitting or other requirements of federal or state
underground injection control programs. Withdrawal
and treatment approaches may be subject to
regulation under federal or state air pollution control
programs or to pretreatment requirements if
contaminated ground water will be discharged to a
municipal wastewater treatment system. Also,
pumping from an aquifer may involve a state's
ground-water regulations on well construction
standards and well spacing requirements as well as
interfere with various competing legal rights to pump
ground water.
Other factors influencing remediation decisions are
the availability of alternate sources of water supply,
the political and judicial pressure, and the availability
of funds. If alternate water supplies are plentiful and
economical, there may be little incentive for more
than cosmetic remediation, if any. Conversely, if there
is great pressure from the public, press and/or courts
to "do something", there is a tendency to overreact-
-to install remediation measures that offer more in
appearance than in substance. In the final analysis,
responsible agencies can pursue only those
remediation measures for which they have resources.
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3.8 References
Borden, R.C., M.D. Lee, J.T. Wilson, C.H. Ward and
P.B. Bedient. 1984. Modeling the Migration and
Biodegradation of Hydrocarbons Derived from a
Wood-Creosoting Process Waste. Proceedings of
Petroleum Hydrocarbons and Organic Chemicals in
Ground Water: Prevention, Detection, and
Restoration, Conference, November 5-7, 1984,
Houston, ~TX.
Bouwer, E.J., and P.L. McCarty. 1983.
Transformation of Halogenated Organic Compounds
Under Dentrification Conditions. Applied and
Environmental Microbiology 45(4)1295-1299.
Clapp, R.B., and G.M. Horn berger. 1978. Empirical
Equations for Some Soil Hydraulic Properties. Water
Resources Research 14:601-604.
Cooper, D.G. 1982. Biosurfactants and Enhanced Oil
Recovery. Proceedings of the 1982 International
Conference on Microbiological Enhancement of Oil
Recovery, May 16-21, 1982, Shangri-La, Afton,
OK.
Cooper, D.G., and J.T. Zajic. 1980. Surface-Active
Compounds from Microorganisms. Advanced Applied
Microbiology 26:229-253.
De Pastrorich, T.L., Y. Baradat, R. Barthal, A.
Chiarelli, and D.R. Fussel. 1979. Protection of Ground
Water from Oil Pollution. CONCAWE Report No.
3179, The Oil Companies' International Study Group,
Den Haag, The Netherlands.
Ehrenfeld, J., and J. Bass. 1984. Evaluation of
Remedial Action Unit Operations of Hazardous Waste
Disposal Sites. Pollution Technology Review No. 110.
Noyes Publications, Park Ridge, NJ.
Ehrlich, G.G., R.A. Schroeder, and P. Martin. 1985.
Microbial Populations in a Jet-Fuel Contaminated
Shallow Aquifer at Tustin, California. U.S. Geological
Survey Open File Report 85-335.
Henson, R.W. and R.E. Kallio. 1957. Inability of
Nitrate to Serve as a Terminal Oxidant for
Hydrocarbons. Science 125:1198-1199.
Jones, J.N., R.M. Bricka, T.E. Myers, and D.W.
Thompson. 1985. Factors Affecting
Solidification/Stabilization of Hazardous Waste.
Proceedings of the Eleventh Annual Research
Symposium for Land Disposal of Hazardous Waste.
EPA-600/9-85-013, U.S. Environmental Protection
Agency, Hazardous Waste Environmental Research
Laboratory, Cincinnati, OH.
Knox, R.C., L.W. Canter, D.F. Kincannon, E.L. Stover
and C.H. Ward. 1984. State-of-tne-Art of Aquifer
Restoration. EPA-600/2-84/182a and b, U.S.
Environmental Protection Agency, Robert S. Kerr
Environmental Research Laboratory, Ada, OK.
Kuhn, E.P., P.J. Colberg, J.L. Schoor, D. Wanner,
A.J.B. Zehnder, and R.P. Schwarzenbach. 1985.
Environmental Science and Technology 19:961-968.
Lee, M.D., and C.H. Ward. 1985. Restoration
Techniques for Aquifers Contaminated with
Hazardous Waste. Journal of Hazardous Materials (In
Press).
Lee, M.D., and C.H. Ward. 1984. Reclamation of
Contaminated Aquifers: Biological Techniques.
Proceedings of the 1984 Hazardous Material Spills
Conference. April 9-12, 1984, Nashville, TN.
Overcash, M.R., and D. Pal. 1979. Design of Land
Treatment Systems for Industrial Waste - Theory
and Practice. Ann Arbor Science. Ann Arbor, Ml.
Parsons, F., G.B. Lage, and R. Rice. 1985.
Biotransformation of Chlorinated Organic Solvents in
Static Microcosms. Environmental Toxicology and
Chemistry 4:739-742.
Parsons, F., P.R. Wood, and J. DeMarco. 1984.
Transformations of Tetrachloroethene and
Trichlonethene in Microcosms and Ground Water.
Journal American Water Works Association
76(2):56-59.
Perry, J.J. 1979. Microbial Cooxidations Involving
Hydrocarbons. Microbiology Review 43:59-72.
Raymond, R.L. 1974. Reclamation of Hydrocarbon
Contaminated Ground Waters. U.S. Patent Office,
3,846,290. Patented November 5, 1974.
Suflita, J.M., and S.A. Gibson. 1985. Biodegradation
of Haloaromatic Substrates in a Shallow Anoxic
Ground Water Aquifer. Proceedings of the Second
International Conference on Ground Water Quality
Research, March 26-29, 1984, Tulsa, OK.
Suflita, J.M., and G.D. Miller. 1985. Microbial
Metabolism of Chlorophenolic Compounds in Ground
Water Aquifers. Environmental Toxicology and
Chemistry 4:751-758.
U.S. Army Engineers. 1985. Guidelines for
Preliminary Selection of Remedial Actions for
Hazardous Waste Sites. EM 1110-2-505, DA-
USAE, Washington, DC.
U.S. Environmental Protection Agency. 1985.
Remedial Action at Waste Disposal Sites.
EPA/625/6-82-006, U.S. Environmental Protection
Agency, Hazardous Waste Environmental Research
Laboratory, Cincinnati, OH.
van der Waarden, M., L.A. Bridie, and W.M.
Groenewoud. 1977. Transport of Mineral Oil
Components to Ground Water II. Water Research
11:359-365.
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Wilson, B. 1985. Behavior of Trichloroethylene, 1,1-
Dichloroethylene in Anoxic Subsurface Environments.
M.S. Thesis, University of Oklahoma.
Wilson, B.H., and J.F. Rees. 1985. Biotransformation
of Gasoline Hydrocarbons in Methanogenic Aquifer
Material. Proceedings of the NWWA/API Conference
on Petroleum Hydrocarbons and Organic Chemicals
in Ground Water, November 13-15, 1985, Houston,
TX.
Wilson, J.L and S.H. Conrad. 1984. Is Physical
Displacement of Residual Hydrocarbons a Realistic
Possibility in Aquifer Restoration? Proceedings of the
Petroleum Hydrocarbons and Organic Chemicals in
Ground Water: Prevention, Detection, and Restoration
Conference, November 5-7, 1984, Houston, TX.
Wood, P.R., R.F. Lang, and I.L. Payan. 1985.
Anaerobic Transformation, Transport, and Removal of
Volatile Chlorinated Organics in Ground Water. In:
Ground Water Quality, edited by C.H. Ward, W. Giger
and P.L. McCarty, John Wiley & Sons, New York, NY.
Young, L.Y. 1984. Anaerobic Degradation of Aromatic
Compounds. In: Microbial Degradation of Aromatic
Compounds, edited by D.R. Gibson, Marcel Dekker,
New York, NY.
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CHAPTER 4
BASIC HYDROGEOLOGY
Hydrogeology is the study of ground water -- its
origin, occurrence, movement, and quality. Ground
water is a part of the hydrologic cycle and, in order to
understand the influence of the hydrologic cycle on
ground water, it is essential to have some basic
knowledge of precipitation, infiltration, the relationship
between ground water and streams, and the impact of
the geologic framework on water resources. This
chapter provides a brief outline of these topics.
4.1 Precipitation
Much precipitation never reaches the ground; it
evaporates in the air and from trees and buildings.
That which reaches the land surface is variable in
time, area! extent, and intensity. The variability has a
direct impact on streamflow, evaporation,
transpiration, soil moisture, ground-water recharge,
ground water, and ground-water quality. Therefore,
precipitation should be examined first in any type of
hydrogeologic study in order to determine how much
is available, its probable distribution, and when and
under what conditions it is most likely to occur. In
addition, a determination of the amount of
precipitation is the first step in a water balance
calculation.
4.1.1 Seasonal Variations in Precipitation
Throughout much of the United States, the spring
months are most likely to be the wettest months. This
is because low intensity rains often continue for
several days at a time. The rain, in combination with
springtime snowmelt, will saturate the soil and
streamflow is generally at its peak over a period of
several weeks or months. Because the soil is
saturated, this is the major period of ground-water
recharge. In addition, because all of the surface runoff
consists of precipitation and snowmelt, surface waters
most likely will contain less dissolved mineral matter
than at any other time during the year.
Not uncommonly, the fall is also a wet period
although precipitation is not as great or prolonged as
in the spring. Because ground-water recharge can
occur over wide areas during spring and fall, one
should expect some natural changes in the chemical
quality of ground water in surficial or shallow aquifers.
During the winter in northern states the ground is
frozen, largely prohibiting infiltration and ground-
water recharge. An early spring flow coupled with
widespread precipitation may lead to severe flooding
over large areas.
Summer precipitation is more likely to be convective
in nature and the result is high intensity rainfall that
occurs during a short time interval in a small area.
Most of the rain does not infiltrate, there is a soil-
moisture deficiency, and ground-water recharge
over wide areas is not to be expected. On the other
hand, these typically small, local showers can have a
significant impact on shallow ground-water quality
because some of the water flows quickly through
fractures or other macropores in the unsaturated
zone, carrying water-soluble compounds leached
from the dry soil to the water table. In this case,
certain chemical constituents, and perhaps microbes
as well, may increase dramatically.
4.7.2 Types of Precipitation
Precipitation is classified by the conditions that
produce the rising column of unsaturated air which is
antecedent to precipitation.
Convectional precipitation is the result of uneven
heating of the ground, which causes the air to rise
and expand, vapor to condense, and precipitation to
occur. This is the major type of precipitation during
the summer, producing high intensity, short duration
storms usually of small areal extent. They often cause
flash floods in small basins. Ground-water recharge
caused by convective storms is likely to be of a local
nature.
Orographic precipitation is caused by topographic
barriers that force the moisture laden air to rise and
cool. This occurs, for example, in the Pacific
Northwest where precipitation exceeds 100 inches
per year, and in Bangladesh, which receives more
than 425 inches per year nearly all of which falls
during the monsoon season. In this vast alluvial plain,
rainfall commonly averages 106 inches during June
for a daily average exceeding 3.5 inches.
Cyclonic precipitation is related to large low pressure
systems that require 5 or 6 days to cross the United
States from the northwest or Gulf of Mexico. These
51
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systems are the major source of winter precipitation.
During the spring, summer, and fall they lead to rainy
periods that may last 2 or 3 days or more. They are
characterized by low intensity and long duration, and
cover a wide area. They probably have a major
impact on natural recharge to ground-water systems
during the summer and fall and impact ground-water
quality as well.
4.7.3 Recording Precipitation
The amount of precipitation is measured by recording
and nonrecording rain gages. Many are located
throughout the country but, because of the
inadequate density of gages, our estimate of annual,
and particularly summer, precipitation is too low.
Records can be obtained from the Climatological
Data, which are published annually by The National
Oceanic and Atmospheric Administration (NOAA).
Precipitation is highly variable, both in time and
space. The area! extent of precipitation is evaluated
by means of contour or isohyet maps (Figure 4-1).
A rain gage should be installed in the vicinity of a site
under investigation in order to know exactly when
precipitation occurred, how much fell, and its
intensity. Data such as these are essential to the
interpretation of hydrographs of both wells and
streams, and they provide considerable insight into
the causes of fluctuations in ground-water quality.
4.2 Infiltration
The variability of streamflow depends on the source
of the supply. If the source of streamflow is from
surface runoff, the stream will be characterized by
short periods of high flow with long periods of low
flow or no flow at all. Streams of this type are known
as "flashy." If the basin is permeable, there will be
little surface runoff and ground water will provide the
stream with a high sustained uniform flow. These
streams are known as "steady." Whether a stream is
steady or flashy depends on the infiltration of
precipitation and snowmelt.
When it rains, some of the water is intercepted by
trees or buildings, some is held in low places on the
ground (this is known as depression storage), some
flows over the land surface without infiltrating and
eventually reaches a stream (surface runoff), some is
evaporated, and some infiltrates. Of the water that
infiltrates, some replenishes the soil-moisture
deficiency, if any, while the remainder percolates
deeper, perhaps becoming ground water. The
depletion of the soil-moisture begins immediately
after a rain due to evaporation and transpiration.
Infiltration capacity (f) is the maximum rate at which a
soil is capable of absorbing water in a given condition.
Several factors control infiltration capacity:
o Antecedent rainfall and soil-moisture conditions.
Soil moisture fluctuates seasonally, usually being
high during winter and spring and low during the
summer and fall. If the soil is dry, wetting the top
of it will create a strong capillary potential just
under the surface, supplementing gravity. When
wetted, the clays forming the soil swell, which
reduces the infiltration capacity shortly after a rain
starts.
o Compaction of the soil due to rain.
o Inwash of fine material into soil openings, which
reduces infiltration capacity. This is especially
important if the soil is dry.
o Compaction of the soil due to animals, roads,
trails, urban development, etc.
Figure 4-1 Distribution of annual average precipitation in Oklahoma, 1970-79 (from Pettyjohn and others).
16 18 20 22 24 26 28 30 32 34 36 38404244 A
16
18 20
32 —-^ Lines of equal precipitation ^
(inches)
'->, Lir
I (in
24
48
52
52
-------
o Certain microstructures in the soil will promote
infiltration, such as openings caused by burrowing
animals, insects, decaying roots and other
vegetative matter, frost heaving, dessication
cracks, and other macropores.
o Vegetative cover, which tends to increase
infiltration because it promotes populations of
burrowing organisms and retards surface runoff,
erosion, and compaction by raindrops.
o Decreasing temperature, which increases water
viscosity, reducing infiltration.
o Entrapped air in the unsaturated zone, which
tends to reduce infiltration.
o Surface gradient.
Infiltration capacity is usually greater at the start of a
rain that follows a dry period, but it decreases rapidly
(Figure 4-2). After several hours it is nearly constant
because the soil becomes clogged by particles and
swelling clays. Thus a sandy soil, as opposed to a
clay-rich soil, may maintain a high infiltration
capacity for a considerable time.
Figure 4-2 Infiltration capacity decreases with time
during a rainfall event.
Coarse Texture
Fine Texture ,
1 2
Time (hours)
As the duration of rainfall increases, infiltration
capacity continues to decrease. This is partly due to
the increasing resistance to flow as the moisture front
moves downward; that resistance is a result of
frictional increases due to the increasing length of
flow channels and the general decrease in
permeability owing to swelling clays. If precipitation is
greater than the infiltration capacity, surface runoff
occurs. If precipitation is less than the infiltration
capacity, all moisture is absorbed.
When a soil has been saturated by water, then
allowed to drain by gravity, the soil is said to be
holding its field capacity of water. Drainage generally
requires no more than two or three days and most
occurs within one day. A sandy soil has a low field
capacity that is reached quickly; clay-rich soils are
characterized by a high field capacity that is reached
slowly (Figure 4-3).
Figure 4-3 Relation between grain size and field capacity
and wilting point (from Smith and Rune, 1955).
Average Inches Depth of Water Per Foot
Depth of Soil in Plant Root Zone
o
The water that moves down becomes ground-water
recharge. Since recharge occurs even when field
capacity is not reached, there must be a rapid
transfer of water through the unsaturated zone. This
probably occurs through macropores (Pettyjohn,
1983).
4.3 Surface Water
Streamflow, runoff, discharge, and yield of drainage
basin are all nearly synonymous terms. Channel
storage refers to all of the water contained at any
instant within the permanent stream channel. Runoff
includes all of the water in a stream channel flowing
past a cross section; this water may consist of
precipitation that falls directly into the channel,
surface runoff, ground-water runoff, and effluent.
Although the total quantity of precipitation that falls
directly into the channel may be large, it is quite small
in comparison to the total flow. Surface runoff,
including interflow or stormflow, is the only source of
water in ephemeral streams and intermittent streams
during part of the year. It is the major cause of
flooding. During dry weather ground-water runoff
may account for a stream's entire flow. It is the major
53
-------
source of water to streams from late summer to
winter; at this time streams are also most highly
mineralized under natural conditions. Ground water
moves slowly to the stream, depending on the
hydraulic gradient and permeability; the contribution is
slow but the supply is steady. When ground-water
runoff provides a stream's entire discharge, the flow
is called dry-weather flow. Other sources of runoff
include the discharge of industrial or municipal
effluent, or irrigation return flow.
4.3.1 Stream Types
Streams are generally classified on the basis of their
length, size of the drainage basin, or discharge; the
latter is probably the most significant index of a
stream's utility in a productive society. Rates of flow
are generally reported as cubic feet per second (cfs),
millions of gallons per day (mgd), acre-feet per day,
month, or year, cfs per square mile of drainage basin
(cfs/mj2), or inches depth on drainage basin per day,
month, or year. In the United States, the most
common unit of measurement for rate of flow is cubic
feet per second (cfs). The discharge (Q) is
determined by measuring the cross-sectional area of
the channel (A), in square feet, and the average
velocity of the water (v), in feet per second, so that:
Q = vA
(4-1)
From a hydrogeologic point of view, there are three
major stream types -- ephemeral, intermittent, and
perennial. They are determined by the relation
between the water table and the stream channel.
An ephemeral stream owes its entire flow to surface
runoff, it may have no well-defined channel, and the
Figure 4-4 Relation between water table and stream type.
water table consistently remains below the bottom of
the channel (Figure 4-4). Water leaks from the
channel into the ground, recharging the underlying
strata.
Intermittent streams flow only part of the year,
generally from spring to midsummer, as well as
during wet periods. During dry weather these streams
flow only because of the ground water that
discharges into them. This is possible because the
water table is then above the base of the channel
(Figure 4-4). Eventually sufficient ground water has
discharged throughout the basin to lower the water
table below the channel, which then becomes dry.
This reflects a decrease in the quantity of ground
water in storage. During late summer or fall, a wet
period may temporarily cause the water table to rise
enough for ground water to again discharge into the
stream. Thus during part of the year the flood plain
materials are full to overflowing, which causes the
discharge to increase in a downstream direction, but
at other times water will leak into the ground, causing
a reduction of the discharge.
Many streams, particularly those in humid and
semiarid regions, flow throughout the year. These are
called perennial streams. In these cases, the water
table annually remains above the stream bottom,
ground water is discharged, and streamflow increases
downstream (Figure 4-4). A stream in which the
discharge increases downstream, is called a gaining
stream. When the discharge of a stream decreases
downstream due to leakage, it is called a losing
stream.
Losing Stream
(A-A'l
Gaining Stream in Spring
Losing Stream in Fall
(B-B'l
Gaining Stream
(C-C-)
Water Table
in Spring (S)
in Fall (F)
54
-------
Figure 4-5 Water quality data for Cottonwood Creek near Navina, Oklahoma (from U.S. Geological Survey Water
Resources Data for Oklahoma).
ARKANSAS Rivrn RASIM
07159720 COTlOHWnoO CHFFK NEAR t4AV!HA, (X—Continued
WAIER-OUADTY nrcnnos
PERIOD OF RECORD.—Water years 1978 to current year.
PERIOD OF DAILY RECORD.—
SPECIFIC CONDUCTANCE: October 1977 to November 1980.
WATER TEMPERATURE) October 1977 to November 1980.
REMARKS.—Samples *ere collected monthly and specific conductance, pH, water temperature, and dissolved oxygen were
determined In the field.
WATER QUALITY DATA, WATER YEAR OCTOOER 1982 TO SEPTEMBER 1983
DATE
OCT
27...
NOV
29...
DEC
15...
3AN
18...
FEH
2J...
HAR
29...
APR
27...
HAY
21...
JUM
21...
SEP
15...
DATE
OCT
27...
NOV
29...
DEC
15...
JAN
18...
FEB
23...
HAR
29...
APR
27...
HAY
24...
3UN
21...
SEP
15...
TIME
1330
1300
1320
1300
13*5
1*30
1*30
1330
12*5
1030
MAGNE-
SIUM,
DIS-
SOLVED
(MC/L
AS HC)
30
24
36
39
30
3(
40
26
44
29
AGENCY SPE-
ANA- STREAM- CIF 1C
LY7.IHG FLOW, COM-
SAHPLE INSTAH- DtlCT-
(COOE TAHEOUS ANTE
NUMBER) (CFS) (UMHOS)
80020
80020
80020
80020
80020
80070
80020
80020
80020
80020
WATER
SODIUM,
DIS-
SOLVED
15
55
28
26
100
13*
76
266
57
2*
QUALITY
(MT./L PERCENT
AS HA)
HO
8»
130
150
77
92
110
62
120
140
SODIUM
»5
39
*1
44
3*
34
37
33
36
46
1*00
935
1300
1*30
955
1100
1290
850
1300
1320
DATA,
SODIUM
AD-
SOflP-
UOH
RATIO
3
2
3
)
Z
2
2
2
3
J
PH
(STAND-
ARD
UNITS)
7.7
8.0
7.8
7.9
7.7
7.6
7.8
7.8
7.6
7.7
WATER YEAR
POTAS-
SIUM,
DTS-
SOLVFD
(MG/L
AS K)
9.5
5.8
7.7
8.9
4.8
4.4
5.2
4.3
5.9
11
TEMPER -
Arurc
(DEC C)
14.5
7.0
6.0
5.0
10.0
8.0
19.0
20.0
?*.o
21.5
OCTOBER
ALKA-
LINITY
LAO
(MC/L
AS
CAC03)
229
173
2*1
28*
191
213
2*7
188
289
206
OXYCEM,
OIS-
SOLVFD
(MC/L)
£.8
8.2
8.0
6.6
7.4
8.2
4.9
6.3
5.*
6.)
1982 TO
SULFATE
OIS-
OXYGEN,
DIS-
SOLVED
(PER-
CENT
SATUR-
ATION)
66
71
66
55
69
72
55
78
67
75
SEPTEMBER
CHLO-
RIDE,
DIS-
SOI.VFD SOLVFD
(HC/L
AS SO*)
220
160
230
250
170
220
230
1*0
2*0
2*0
(HR/L
AS CD
170
95
150
170
95
100
1*0
71
130
170
HARD-
NT S3
(MC/L
AS
CAC03)
360
280
400
410
310
380
410
270
460
350
1983
SILICA,
DIS-
SOLVED
(MG/L
AS
SI02)
13
10
14
11
12
12
12
12
17
12
HARD-
NESS
NONCAR-
BONATE
(MG/L
AS
CAC03)
127
108
157
127
123
170
161
87
168
141
SOLIDS,
RESIDUE
AT 180
DEC. C
DIS-
SOLVED
(MC/L)
850
574
865
895
591
718
789
508
870
852
CALCIUM
DIS-
SOLVED
(HC/L
AS CA)
93
73
100
100
76
94
97
67
110
91
SOLIDS,
SUM OF
CONSTI-
TUENTS,
DIS-
SOLVED
(MC/L)
810
560
810
900
580
690
780
500
840
820
55
-------
Nearly all water courses have headwater regions
characterized by ephemeral streams. Farther
downbasin, intermittent streams predominate and,
even farther, the water courses are perennial. Some
streams fed by springs or glacial meltwater are
perennial throughout their entire length.
The natural gradation from one stream type to
another may be interrupted by either natural or man-
made causes. Irrigation may provide enough recharge
to cause the water table to rise sufficiently to increase
ground-water runoff, while pumping from wells may
have the opposite effect.
Streams flowing through saturated permeable
deposits, such as sand and gravel, are normally
gaining streams, but streams flowing through karst
regions may be losing in one reach and gaining in
another. High dry-weather flow may reflect the
discharge of water from mine workings.
From a hydraulic perspective, a stream is similar to
an exceedingly long, very shallow, horizontal well.
Consequently, the chemical quality of water in the
stream during dry weather reflects the quality of
ground water in the zone of active circulation within
the basin if the stream is not contaminated by some
surface source. During wet weather, the chemical
quality of water in a stream varies largely because of
the mixing of dilute surface runoff with the more
highly mineralized ground-water runoff (Figure 4-5).
The sediment load, reflecting erosion in the basin and
stream channel, also affects the quality of the stream.
The loading of a stream with either sediment or
dissolved constituents is commonly reported in units
of tons per day (Tons per day = Discharge x
Concentration x .0027).
4.3.2 Stream Discharge Measurements and
Records
At a stream gaging site the discharge is measured
periodically at different rates of flow, which are plotted
against the elevation of the water level in the stream
(stage or gage-height). This forms a rating curve
(Figure 4-6). At a gaging station the stage is
continuously measured and this record is converted,
by means of the rating curve, into a discharge
hydrograph. The terminology used to describe the
various parts of a stream hydrograph are shown in
Figure 4-7.
Discharge, water quality, and ground-water level
records are published each year by the U.S.
Geological Survey for each state. An example of the
annual record of a stream is shown in Figure 4-8.
Notice that these data are reported in "water years."
The water year is designated by the calendar year in
which it ends, which includes 9 of the 12 months.
Thus, water year 1985 extends from October 1, 1984
to September 30, 1985.
Figure 4-6 A generalized stream stage vs. discharge
rating curve.
o>
'5
i i i i
Discharge (cubic feel/second)
Figure 4-7 Stream hydrograph showing definition of
terms.
Crest
Time (days)
4.4 The Relation Between Surface Water
and Ground Water
There are many tools for learning about ground water
without basing estimates on the ground-water
system itself - that is, one can use streamflow
data. Analyses of streamflow data permit an
evaluation of the basin geology, permeability, the
amount of ground-water contribution, and the major
areas of discharge. In addition, if chemical quality
data are available or collected for a specific stream,
they can be used to determine background
56
-------
Figure 4-8 Stream discharge record for Cottonwood Creek near Navina, Oklahoma (from U.S. Geological Survey Water
Resources Data for Oklahoma).
07159720
ARKANSAS RIVFR BASIN
COUONWOOO CRFEK NT.AR NAVINA, OK
LOCATION.--Lat )5*46')6", lonq 97')2'45", SW 1/4 MW 1/4 srr.. 17, 1.15 M., R.4 W., Logan County, llydrologtc Unit
11050002 on downstream right hank, 0.5 ml (0.8 km) downstream from Oner Creek, 1.7 ml (2.1 km) southeast of
Navina, 10.7 ml (17.2 km) southwest of Outhrle, and at mile 25.0 (40.2 km).
DRAINAGE AREA.— 2*7 ml2 (6*0 km2).
WATER-DISCHARGE RECORDS
PERIOD OF RECORD.—October 1977 to September 1980, March 1982 to current year.
CAGE.—Water-stage recorder. Datum of gaqe Is 962.10 ft (293.248 m) National Ceort>tlc Vertical Datum of 1929.
REMARKS.—Records poor. Low flow sustained by part of sewage effluent from Oklahoma City.
EXTREMES FOR PERIOD OF RECORD
(6.8)7 m)| minimum dally
EXTREMES FOR CURRENT YEAR.--Maximum
m), no other peak above base
22, 2), 26.
.--Maximum discharge, 12,300 ft3/s (348 m'/») May 30, 1980, gage height, 22.43 ft
, 8.0 ft'/s (0.2) mVs) Oct. 14, 15, 1977.
m discharge. 3,600 ft*/» (10? m'/») at 06*5 May 14, gage height, 20.87 ft (6.361
of 2,000 ft'/s (56.6 «T/s)l minimum dally discharge, 15 ft /s (0.42 m /s) Oct.
DAY
OCT
Discharge, In Cubfc Feet per Second, Water Year October 1982 to September 1983
Mean Values
NOV
DEC
3AN
FEB
MAR
APR
MAY
3UN
DDL
WTR YR 198)
TOTAL
)1281
MEAN
85.7
MAX
3330
MIN
15
AC-FT 62050
AUG
SEP
1
2
)
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
TOTAL
MEAN
MAX
MIN
AC-FT
31
31
36
31
30
28
32
26
24
21
21
22
22
20
3)
34
24
18
31
19
18
15
15
16
17
15
16
24
26
21
27
744
24.0
36
15
1480
32
36
34
37
39
41
44
46
49
47
48
55
46
41
46
40
40
38
38
39
43
)!>
38
37
36
40
68
73
52
38
1300
43.3
73
32
2580
27
25
24
32
42
34
28
24
23
21
30
29
26
26
27
22
22
23
25
23
21
?.z
21
25
40
31
29
58
45
35
32
892
28.8
58
21
1770
27
26
24
24
25
24
24
22
24
28
28
28
20
16
22
23
26
26
21
29
28
28
31
34
30
32
65
41
36
33
34
879
28.4
65
16
1740
199
186
93
72
57
57
52
48
50
80
68
52
49
48
47
45
47
44
42
68
99
210
101
68
58
56
51
48
—
—
2095
74.8
210
42
4160
47
48
51
52
67
66
58
51
45
43
41
41
44
48
48
48
48
56
48
44
48
46
45
44
47
49)
478
197
143
123
110
2768
89.)
49)
41
5490
105
103
94
92
196
302
164
126
117
150
142
114
101
93
85
82
82
80
77
77
75
8?
274
205
105
85
75
68
63
63
...
3477
116
302
63
6900
64
62
58
56
5)
52
50
46
46
48
447
146
942
33)0
1980
467
294
3?6
27)
705
721
550
389
267
189
156
140
131
123
134
393
12140
392
33)0
46
24080
217
159
1)8
118
107
108
94
88
82
78
329
259
243
199
96
SO
71
67
64
60
54
52
48
47
49
63
81
1)9
465
150
—
3805
177
465
47
7550
86
6)
53
45
42
40
38
37
35
)5
)2
32
)1
29
)0
29
)0
)1
)0
28
27
26
25
24
2)
2)
21
20
20
23
21
1029
)).2
96
20
2040
21
20
19
19
19
19
21
2)
24
22
21
20
20
21
20
19
19
19
19
178
476
142
81
55
46
40
)6
))
30
28
26
1506
48.6
476
19
2990
24
24
22
22
21
22
20
20
19
20
20
19
21
26
22
19
18
19
18
25
48
77
70
19
21
20
19
20
18
18
646
21.5
48
18
1280
57
-------
concentrations of various parameters and locate
areas of ground-water contamination as well.
The interrelationship between surface and ground
water is of great importance in both regional and local
hydrologic situations, and a wide variety of
information can be obtained by analyzing stream flow
data. Evaluation of the ground-water component of
streamflow can provide important and useful
information regarding regional recharge rates, aquifer
characteristics, ground-water quality, and indicate
areas of high potential yield to wells. To determine
the ground-water component of runoff a stream
hydrograph must first be separated into its component
parts. There are many ways in which this can be
accomplished, although all are quite subjective.
Several methods will be briefly described.
4.4.7 The Regional System
In order to better appreciate the origin and
significance of ground-water runoff and its quality,
one should briefly examine the regional ground-
water flow system. In humid and semiarid regions, in
particular, the water table generally conforms to the
surface topography. The hydraulic gradient or water
table slopes away from divides and topographically
high areas toward adjacent low areas, such as
streams and rivers. The high areas serve as ground-
water recharge areas, while the low places are
ground-water discharge zones (Figure 4-9).
As water infiltrates in a recharge area, the mineral
content is relatively low. The quality changes,
however, along the flow path and dissolved solids as
well as other constituents increase with increasing
distances traveled in the ground. The water eventually
flows into a stream or body of surface water and, due
to the different lengths of flow paths and rock
solubility, even streams and small lakes in close
proximity may have large differences in both flow and
quality.
4.4.2 Bank Storage
As a flood wave passes a particular stream cross
section, the water table may rise in the adjacent
stream side deposits. This occurs because the
elevation of the water level in the stream, or the
"stage," rises quickly and soon becomes higher than
the water table. This blocks the ground water that
would normally flow into the stream and causes the
water table to rise in the flood plain. In addition,
because of the higher stage, water will flow from the
stream into the ground. Once the stage begins to fall,
the water, which was recently added to the ground
Figure 4-9 The chemical quality of ground water commonly changes along a flow path in the regional system as water
flows from areas of recharge to areas of discharge.
Recharge Area
Head decreases and dissolved solids
increase with depth
Discharge Area
Recharge Area
Head increases and dissolved solids
decrease with depth
Head decreases and dissolved solids
increase with depth
58
-------
water, will begin to flow back into the stream, rapidly
at first and then more slowly as the water-table
gradient declines. This temporary storage of water in
the near vicinity of the stream channel is called bank
storage (Figure 4-10).
As the drainage from bank storage progresses, the
recession segment of the hydrograph gradually tapers
off into what is called a depletion curve, the shape of
which is controlled by the permeability of the stream
side deposits. A master depletion curve is used to
separate a stream hydrograph.
4.4.3 Master Depletion Curve
Intervals between surface runoff events are generally
short and, therefore, depletion curves must be
constructed from a combination of several arcs of the
hydrograph with the arcs overlapping in their lower
parts (Figure 4-11). To plot a depletion curve,
tracing paper is placed over a hydrograph of daily
flows and, using the same horizontal and vertical
scales, the lower arcs are traced, working backward
in time from the lowest discharge to a period of
surface runoff. The tracing paper is moved
horizontally until the arc of another runoff event
coincides in its lower part with the arc already traced.
The process is continued until all the available arcs
are plotted on top of one another. The upward
curving parts of individual arcs are disregarded. The
resulting continuous arc is a mean or normal
depletion curve that presumably represents the
hydrograph that would result from the ground-water
runoff alone during a prolonged dry period.
4.4.4 Separating a Hydrograph by Graphical
Methods
A hydrograph can be separated in the following ways.
A depletion curve is positioned on the lower part of
the recession limb of a runoff hydrograph, as shown
in Figure 4-12. Notice that it departs from the actual
recession curve at point D, which should reflect the
end of surface runoff. The master curve is extended
backward to its intersection at C with a vertical line
drawn through the peak. A second line originating at
A, which is the start of surface runoff, is drawn to C.
The area or discharge below the line ACD is
ground-water runoff.
It may be difficult to locate D with the depletion curve
and a second method is to estimate its position with
the equation:
N = A2 (4-2)
where:
N = number of days after a peak when surface
runoff ceases
A = drainage basin area, in square miles.
The distance N can be measured directly on the
hydrograph.
Another method for separating a hydrograph consists
of extending a line from point A, the start of surface
runoff, to point D (Figure 4-12). A third method
consists of extending the presurface runoff depletion
trend to a point directly under the hydrograph peak,
B, and then from B to D. This reflects a stream that is
influenced by bank storage.
4.4.5 Separating a Hydrograph by Chemical
Methods
Hydrographs also can be separated by chemical
means. During baseflow the natural quality of a
stream is at or near its maximum concentration of
dissolved solids but, as surface runoff reaches the
channel and provides an increasing percentage of the
flow, the mineral concentration decreases. After the
peak, ground-water runoff increases, surface runoff
decreases, and the mineral content increases.
Several investigators (including Toler, 1965; Kunkle,
1965; Pinder and Jones, 1969; Visocky, 1970; and
LaSala, 1968) have used the relation between runoff
and water quality to calculate ground-water runoff
from one or more aquifers or to measure streamflow.
This method is based on the concentration of a
selected chemical parameter that is characteristic of
ground-water and surface runoff. The basic
equation, which can take several forms, is as follows:
Qg = Q (C - Cs)/(Cg - Cs)
where:
Qg = quantity of ground-water runoff
C = concentration of the specific chemical
parameter on conductance of runoff
Q = runoff
Cs = concentration of the specific chemical
parameter or conductance of surface runoff
Qs = surface runoff
Cg = concentration of the specific chemical
parameter or conductance of ground water.
Specific conductance is most often used because of
the ease in obtaining it. Cg is measured in a well or
series of Dwells and it should be about the same as C
in a stream during baseflow. Cs is measured from a
sample collected from the surface of the ground
before the water reaches the stream. It is assumed
that Cs and Cg are constant. Q and C are measured
directly in the stream.
Toler (1965) used this method during baseflow to
determine the quantity of water discharging from a
surficial sand aquifer (Qi) and an underlying artesian
limestone aquifer (Q2) in Florida. In this case, as
shown in Figure 4-13, the dissolved solids in water
from the limestone (C2) averaged 50 mg/l, while that
from the sand (Ci) averaged 10. When the stream
had a discharge of 18 cfs (Q) and corresponding
59
-------
Figure 4-10 Movement of water into and out of bank storage along a stream in Indiana, (from Daniels ef a/, 1970),
Land Surface
13
11
ABC D E
I I I I I '
(D
X
ID
01
Peak 3 .
0 150
Began 1700 Hours
13
11
.C
o>
'3
0)
O)
Sand
I I I I
I
I I I
200 400 600 800
Horizontal Distance (feet)
Land Surface
1000
1200
13
11
.c
O>
'5
-------
Figure 4-12 A stream hydrograph can be separated by
three different methods.
Peak
Figure 4-13 Schematic showing the contribution of water
from different aquifers to Econfina Creek,
Florida.
dissolved solids of 43 mg/l (C), ground water from the
limestone was discharging through a series of springs
at a rate of about 14.85 cfs:
Q2 = Q (C - Ci)/(C2 - Ci)
= 18 (43 - 10)/(50 - 10) = 14.85 cfs
Kunkle (1965) used specific conductance
measurements to separate a runoff event hydrograph
of Iowa's Four Mile Creek. In this case, continuous
recordings of discharge and conductance were
available. Specific conductance of the ground water
and the stream at low flow averaged 520 micromhos
(Ci), while surface runoff averaged 160 (62).
Instantaneous ground-water runoff during the event
was calculated for several points under the
hydrograph (Figure 4-14). For example, when the
discharge and conductivity of Four Mile Creek was
2.3 cfs (Q) and 410 micromhos (C), respectively,
ground-water runoff (Qi) was 1.6 cfs:
Ql = Q (C - C2)/(Ci - C2)
= 2.3 (410 - 160)7(520 - 160) = 1.6 cfs
4.4.6 Ground-Water Rating Curves
A ground-water rating curve shows the relation
between the water table and streamflow. Water levels
are measured in one or more wells that are not
influenced by pumping. At the same time, the
discharge is determined during periods of baseflow.
Selected water level and discharge measurements
are then plotted on a graph and a smooth curve is
drawn through the points as illustrated in Figure 4-
15. The rating curve shows what the discharge
should be relative to some particular ground-water
level; the difference, if any, is surface runoff. For
example, in Figure 4-15 when the ground-water
level is 46.4 feet, baseflow should be 20 cfs. If the
stream discharge happened to be 35 cfs, for
example, then the difference, 15 cfs, would have
been caused by surface runoff.
Olmsted and Hely (1962) used a ground-water rating
curve to evaluate the water-bearing properties of
folded igneous and metamorphic rocks in the
Piedmont Upland of the Delaware River Basin. Here
the average depth of water in all of the observation
wells averaged about 17.5 feet and the annual
fluctuation was 5.75 feet; precipitation averages about
44 inches per year. Hydrographs of runoff and
ground-water runoff for Brandywine Creek are
shown in Figure 4-16. The study found that
ground-water runoff accounted for 67 percent of the
total flow over a 6-year period. This compares
favorably with the 64 percent determined for North
Branch Rancocas Creek in the coastal plain of New
Jersey; 74 percent for Beaverdam Creek in the
coastal plain of Maryland (Rasmussen and
Andreasen, 1959); 42 percent for Perkiomen Creek, a
flashy stream in the Triassic Lowland of Pennsylvania;
and 44 percent for the Pomperaug River Basin, a
61
-------
Figure 4-14
1000
500
200
20
10
Hydrographs showing the discharge, specific
conductance, and computed ground-water
runoff in Four Mile Creek, Iowa (from Kunkle,
1965).
Specific Conductance
Discharge Hydrograph
Ground-Water Runoff
Computed from
Conductivity
2223242526272829301 23
September October
small stream in Connecticut (Meinzer and Stearns,
1929).
A single rating curve cannot be used with much
accuracy during certain times of the year when the
water table lies at a shallow depth because of
significant losses of ground water to
evapotranspiration. In their study of Panther Creek in
Illinois, Schicht and Walton (1961) developed two
rating curves, one for use when evapotranspiration is
high, the other when it is low (Figure 4-17). Double
rating curves also can be used to estimate
evapotranspiration losses. For example, in Figure 4-
17 a ground-water level stage for stream of 6 feet
below land surface would indicate about 24 cfs of
ground-water runoff when evapotranspiration is high
and about 48 cfs when it is low. Therefore,
streamflow is depleted by 24 cfs during periods of
high evapotranspiration; this can be converted to
losses per square mile of drainage basin above the
gage.
Various methods of hydrograph separation are
available, all of which are laborious, time consuming,
quite subjective, and open to questions of accuracy
and interpretation. In each case a technique is used
to provide a number of points on a hydrograph
through which a line can be drawn to separate
ground-water runoff from surface runoff. Once this
line is drawn, one must then determine, directly on
the hydrograph, the daily value of each of the
separated components and then sum the results.
4.4.7 Determining Regional Ground-Water
Recharge Rates
Annual ground-water runoff divided by total
discharge provides the percentage of stream flow that
consists of ground water. Effective ground-water
recharge is that quantity of precipitation that
infiltrates, is not removed by evapotranspiration, and
eventually discharges into a stream. It is equivalent to
ground-water runoff.
Effective ground-water recharge rates can be easily
estimated with a computer program (Pettyjohn and
Henning, 1978). This program separates a
hydrograph by three different methods, provides
monthly recharge rates, an annual rate, and produces
a flow-duration curve. The results compare favorably
with those obtained by other means. The data base is
obtained from annual U.S. Geological Survey
streamflow records.
4.4.8 Seepage Measurements
Seepage or dry-weather measurements consist of
discharge determinations made at several locations
along a stream during a short time interval when
runoff is comprised entirely of baseflow. Rather than
actually measuring the discharge, published records
of a single day can be used by merely plotting on a
map the daily mean flow of all the gages in the basin.
Measurements such as these permit an evaluation of
the basin geology, permeability, the amount of
ground-water contribution, and the major areas of
discharge. In addition, if chemical quality data are
collected in the same manner, they can be used to
determine background concentrations of various
parameters and locate areas of ground-water
contamination as well.
The flow of some streams increases substantially
within short distances. Under natural conditions, this
increase probably indicates the presence of deposits
or zones of high permeability adjacent to the stream
channel. These zones may consist of deposits of
sand and gravel, fractures or faults, solution openings
in limestone, or merely local changes in grain size
and increased permeability. In gaining stretches,
62
-------
Figure 4-15 Rating curve of mean ground-water level compared with base flow of Beaverdam Creek, Maryland (from
Rasmussen and Andreason, 1959).
I
• 1950
A 1951
• 1952
Number indicates calendar month; January = 1
J I I I I I I I I I
Base Row (cubic feet/second)
Figure 4-16 Hydrograph of Brandywfne Creek, Chadd's Ford, Pennsylvania, 1952-1953.
5000
Discharge of Brandywine Creek
63
-------
Figure 4-17 Rating curve of mean ground-water level and
base flow in the Panther Creek basin, Illinois
(from Schicht and Walton, 1961).
Periods when
evapotranspiration is
very small
•• Periods when
evapotranspiration is
great
40 80 120 160 200 240
Ground-Water Runoff (cubic feet/second)
280
ground water may discharge through a series of
springs or seeps along valley walls or the stream
channel, or it may seep upward directly into the
channel. During certain periods, particularly
springtime and after heavy rainfalls, ground water
may discharge with such a high velocity that the sand
grains are partly suspended and lose all strength, that
is, they become quicksand.
A number of discharge measurements were made in
the Scioto River Basin, which lies in a glaciated part
of central Ohio. The flow measurements themselves
are important in that they show the actual discharge,
in this case at about the 90 percent flow, which is the
discharge that is equaled or exceeded 90 percent of
the time. In this case the discharge was reported in
units of millions of gallons per day (mgd) since it was
a municipal water-supply study, rather than the
usual cfs. As Figure 4-18 shows, the discharge at
succeeding downstream sites on the Scioto River, as
well as its tributaries, is greater than that immediately
upstream. This shows that the river is gaining and
that water is being added to it by ground-water
runoff from the adjacent deposits.
4.4.9 Maps of Potential Ground-Water Yield
A particularly useful method for evaluating the
hydrogeology of a basin consists of relating the
discharge to the size of the drainage basin (cfs or
mgd per square mile of basin). One can use this
method to examine Figure 4-18 and then relate the
flow index (cfs or mgd/mi2) to the geology and
hydrology of the area. A cursory examination of the
data shows that the flow indices can be conveniently
separated into three distinct but arbitrary groups:
where the flow index is between 0.01 and 0.020
mgd/mi2; between 0.021 and 0.035 mgd/mi2; and
greater than 0.036 mgd/mi2. Notice that even though
several watercourses fall into the larger flow index
group, the actual discharge ranges from 3.07 to 1.81
mgd.
Logs of wells drilled along the streams with a flow
index in the first group show a preponderance of fine
grained material that contains only a few layers of
sand and gravel; these wells generally yield less than
10 gpm (gallons per minute). For the stream
segments in the second group, however, logs of wells
and test holes indicate that several feet of sand and
gravel underlie fine-grained alluvial material, the
latter of which ranges from 5 to about 25 feet in
thickness. Adequately designed and constructed wells
that tap the outwash deposits produce as much as
500 gpm. Glacial outwash, much of it very coarse-
grained, forms an extensive, very permeable deposit
through which flow the streams and river of the third
group. The outwash extends from the surface to
depths that in places exceed 200 feet. Here ground-
water recharge rates range from 200,000 gpd/mi2
during dry months to more than 500,000 in spring.
Industrial wells tapping these deposits can produce
more than 1,000 gpm.
The above example shows that by combining dry-
weather discharge data and well yields with a map
showing the areal extent of the deposits that are
characteristic of the stream valleys, a map can be
developed that indicates the potential yield of the
area. The map of potential ground-water yield relies
heavily on streamflow measurements and some
geologic data, but it provides a good first cut
approximation of ground-water availability.
4.4.10 Quality as an Indicator
Stream reaches typified by significant increases in
ground-water runoff may also have unusual quality.
In northern Ohio the discharge of a small stream that
drains into Lake Erie increases over a 3-mile stretch
from less than 1 to more than 28 cfs and remains
relatively constant afterward. The increase begins at
an area of springs where limestone, which has an
abundance of solution openings, crops out in and
near the stream. The glacial till-limestone surface
dips downstream, eventually exceeding 90 feet in
depth.
In the upper reaches of the stream, baseflow is
provided by ground-water runoff from the adjacent
thin covering of till, which has a low permeability.
Because this water has been in the ground but a
short time, the mineral content is low, as indicated by
the specific conductance of 583 to 638 microohms
(Figure 4-19). Where strearnflow begins to
significantly increase, the limestone aquifer provides
the largest increment. Moreover, the bedrock water
contains excessive concentrations of dissolved solids,
64
-------
Figure 4-18 Discharge and low flow indices of the Scioto River in central Ohio are strongly influenced by local geologic
conditions. These data allow the development of a potential ground-water yield map (from Pettyjohn and
Henning, 1979).
"\
Columbus
167
.0500
Upper number is low flow, mgd.
Lower number is low flow, mgd/sq mi
Area of surficial outwash; well yields
may exceed 1000 gpm.
Area of outwash covered by a few teet
of alluvium; well yields commonly
between 500 and 1000 gpm.
Chillicothe
Generally fine-grained alluvium along
flood plain; well yields usually less than
25 aom.
65
10
I
15
l
20
I
Scale (miles)
-------
Figure 4-19 Fish populations are controlled by discharge of mineralized water from an underlying carbonate aquifer in
Green Creek, in northeastern Ohio.
Station Number
No. species
No. indiv.
D.O.
Q
Temp C
PH
Alk
C02
Cond.
Lake Erie
Glacial Till
10
980
10.8
0
21
8.27
285
4
13
1527
9
0
20.5
8.17
253
1.0
10
256
9.5
11.76
13
7.76
277
14
14
520
11.3
17.59
16.5
7.98
250
6
13
184
9.6
22.28
16
8.07
253
3
13
71
9.0
24.211
17.0
8.13
255
2
2
2
8.7
27.62
17
8.12
638
583
2410
2340
2370
2380
Limestone with Solution Openings
5
hardness, and sulfate and in this stretch calcite
precipitates on rocks in the stream channel. In the
upper reaches of this stream, the fish population is
exceedingly abundant, but in the vicinity of the
springs it diminishes quickly and remains in a reduced
state throughout the remaining length. No doubt the
reduction in fish population is directly related to the
natural quality of the water that flows from the
limestone.
A method to locate relatively small areas of ground-
water contamination by means of stream quality was
described by Pettyjohn (1975, 1985). In this case, the
municipal water supply at the central Ohio city of
Westerville periodically contained excessive
concentrations of chloride, producing a salty taste.
The water was obtained from Alum Creek, which was
being contaminated by oil-field brines from scores of
wells in the 189 square mile upstream part of the
basin. The contamination was largely the result of
leakage of brine from "evaporation pits" to the water
table and, eventually, the contaminated ground water
reached a water course.
In order to locate specific areas of contamination,
water samples were collected from Alum Creek and
many of its small tributaries during a single day in
which the streamflow consisted entirely of ground-
water runoff. The background concentration of
chloride (less than 25 mg/l) was established on the
basis of its concentration in uncontaminated small
tributaries. Concentrations exceeding background
were assumed to be the result of contamination
(Figure 4-20).
The chloride concentrations were plotted on a base
map showing the location of all oil and gas wells and
tests, both operating and abandoned. All
contaminated tributaries contained oil wells and
"evaporation pits" within their subbasins, some of
which were the source of the chloride. Next, the
configuration of each small contaminated basin was
delineated on a topographic map. The well location
provided some control on the point source of
contamination. It was then possible to estimate the
general size of each contaminated site because it had
to lie in the vicinity of a well within that small basin
and the plume had to trend downgradient toward the
stream (Figure 4-21).
The approach described above allows an investigator
to minimize drilling costs for monitoring wells because
uncontaminated areas are readily evident and the
investigator can then key on selected sites. Once
66
-------
Figure 4-20 Distribution of chloride and oil and gas wells in Alum Creek basin, Ohio.
Chloride 25 mg/L or less
—• Chloride >25 <50 mg/L
••Chloride >50 mg/L
• Oil or gas well (existing or abandoned)
• Dry hole (oil or gas test)
A Salt-water disposal well
Sample collection site and number
67
-------
Figure 4-21 Areas of ground-water pollution in Alum Creek basin, Ohio..
Chloride 25 mg/L or less
Chloride >25 <50 mg/L
Chloride >50 mg/L
• Oil or gas well (existing or abandoned)
• Dry hole (oil or gas test)
A Salt-water disposal well
Sample collection site and number
68
-------
contaminated areas have been located, additional
surface water samples can be collected from the
small basins to permit a more detailed assessment.
4.4.11 Temperature as an Indicator
The temperature of shallow ground water is nearly
uniform, reflecting the mean annual temperature of
the region. It ranges from a low of about 37° F in the
north-central part of the United States to more than
77° F in southern Florida (Figure 4-22). Surface
water temperatures, however, range within wide
extremes, freezing in the winter in northern regions
and exceeding 100°F during hot summer days in the
south. Mean monthly stream temperatures during July
and August range from a low of 55 °F in the northwest
to more than 85°F in the southeast (Figure 4-23).
During the summer when ground water provides a
significant increment of flow, the temperature of water
in a stream's gaining reach will decline. Conversely,
during winter the ground water will be warmer than
that on the surface and, although ice will normally
form, parts of a stream may remain open. In central
Iowa, for example, winter air temperatures commonly
drop below zero and ice quickly forms on streams,
ponds, and lakes. Here ground-water temperatures
are about 52°F and, if a sufficient amount is
discharging into a surface water body, ice may not
form. In summer the relatively cold ground water
(52 °F) mixes with the warm (more than 79 °F) surface
water to produce a mixture colder than that in
nongaining reaches.
The point to be made here is that the evaluation of
stream temperature provides clues to changes in
permeability and perhaps even chemical quality.
4.4.12 Flow Duration Curves
A flow-duration curve shows the frequency of
occurrence of various rates of flow. They are useful
for regional evaluations of hydrogeologic conditions.
When used in conjunction with some of the other
methods described above, the investigator can readily
determine areas that are subject to ground-water
contamination. That is, areas and zones that provide,
relatively speaking, large amounts of ground-water
runoff reflect permeable zones that are most sensitive
to contamination.
The flow-duration curve is a cumulative frequency
curve prepared by arranging all discharges of record
in order of magnitude and subdividing them according
to the percentages of time during which specific flows
are equaled or exceeded. All chronologic order is lost
(Cross and Hedges, 1959). Flow-duration curves
may be plotted on either probability or semilog paper.
In either case, the shape of the curve is an index of
the natural storage in a basin, including ground water.
Since dry- weather flow consists entirely of
ground-water runoff, the lower end of the curve
indicates the general hydrogeologic characteristics of
shallow aquifers.
Several flow-duration curves for Ohio streams are
shown in Figure 4-24. During low-flow conditions,
the curves for several of the streams, such as the
Mad, Hocking, and Scioto Rivers as well as Little
Beaver Creek trend toward the horizontal, whereas
Grand River, White Oak and Home Creeks all remain
very steep. The former contain permeable deposits.
Mad River flows through a broad valley that is filled
with very permeable sand and gravel and, as
expected, the river maintains a high sustained flow.
The Hocking River locally contains outwash in and
along its flood plain, which provides a considerable
amount of ground-water runoff. Above Columbus,
the Scioto River flows across thin layers of limestone
that crop out along the stream valley; the adjacent
uplands are covered with glacial till. In this reach,
ground-water runoff is relatively small. Immediately
south of Columbus, however, the valley widens
considerably and is filled with coarse, permeable
outwash. Mad River has a higher low-flow index
than the Scioto River at Chillicothe because the Mad
receives ground-water runoff throughout its entire
length, while the flow of the Scioto increases
significantly only in the area of outwash.
White Oak and Home Creeks originate in bedrock
areas where relatively thin alternating layers of
sandstone, shale, and limestone crop out along the
steep hillsides. The greater relief in these basins
promotes surface runoff and the rocks are
distinguished by moderately low permeability. As the
flow-duration curves indicate, ground-water runoff
from these basins is far less than those that contain
outwash.
The above examples and techniques can be used
mainly for regional hydrogeologic evaluations.
Increases in dry-weather flow, excluding inflow from
tributaries, are usually caused by an increase in
permeability. This, in turn, implies the presence of an
aquifer or zone that might serve as a major source of
water supply and it therefore should be protected.
Abrupt changes in a stream's chemical quality during
dry-weather flow probably will indicate zones of
permeability that are greater than the predominant
strata. The change in quality should indicate the
presence of discharge areas of contaminated
ground-water runoff or the natural chemical quality
of underlying aquifers.
The major purpose of stream hydrograph separation
is to develop an estimate of the amount of ground-
water runoff. If the percentage of ground-water
runoff if large, such as 60 percent or more, then the
rocks within the basin are permeable, infiltration and
ground-water recharge are large, and the basin has
a good potential for the development of ground-
water sources of supply. Consequently, the basin or
69
-------
Figure 4-22 Typical ground-water temperatures (°F) (from Johnson, 1966).
47
67
72
Figure 4-23 Summer stream temperatures (°F).
70
-------
Figure 4-24 Flow-duration curves for selected Ohio streams (from Cross and Hedges, 1959).
10
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Percent of Time Discharge per Square Mile Equalled or Exceeded That Shown
71
-------
Figure 4-25 The water table generally conforms to the surface topography.
Well
Flooded Basement
Stream
Unsaturated Zone
(Soil Moisture)
Saturated Zone
(Ground Water)
Openings Largely
• Filled with Air
Capillary Fringe
Water Table
Openings Filled with
• Water
72
-------
parts of its should be protected because it is readily
subject to contamination.
4.5 Ground Water
The greatest difficulty in working with ground water is
that it is hidden from view, cannot be adequately
tested, and occurs in a complex environment. On the
other hand, the general principles governing ground-
water occurrence, movement, and quality are quite
well known, which permits the investigator to develop
a reasonable degree of confidence in his predictions.
The experienced investigator is well aware, however,
that these predictions are only an estimate of the way
the system functions. Ground-water hydrology is not
an exact science, but it is possible to develop a good
understanding of a particular system if one pays
attention to fundamental principles.
4.5.1 The Water Table
Water under the surface of the ground occurs in two
zones, an upper unsaturated zone and the deeper
saturated zone (Figure 4-25). The boundary
between the two zones is the water table. In the
unsaturated zone, most of the pore space is filled
with air, but water occurs as soil moisture and in a
capillary fringe that extends upward from the water
table. The water in this zone is under a negative
hydraulic pressure; that is, it is less than atmospheric.
Ground water occurs below the water table and all of
the pores are filled with fluid that is under pressure
greater than atmospheric.
The water table conforms to the surface topography,
but it lies at a greater depth under hills than it does
under valleys (Figure 4-25). In general, the water
table lies at depths ranging from 0 to about 20 feet or
so in humid and semiarid regions, but its depth
exceeds hundreds of feet in some desert
environments.
The elevation of the water table must be determined
with care, and many such measurements have been
incorrectly taken. The position of the water table can
be determined from the water level in swamps,
flooded excavations (abandoned gravel pits, highway
borrow pits, etc.), sumps in basements, lakes, ponds,
streams, and shallow dug wells. In some cases there
may be no water table at all or it may be seasonal.
Measurement of the water level in drilled wells,
particularly if they are of various depths, will more
likely reflect the pressure head of one or more
aquifers that are confined than the actual water table.
Accurately determining the position of the water table
is important because the thickness, permeability, and
composition of the unsaturated zone exert a major
control on ground-water recharge and the
movement of contaminants, particu'arly organic
compounds, from land surface to an underlying
aquifer. Attempting to determine the position of the
water table by measuring the water level in drilled
wells nearly always will indicate an unsaturated zone
that is substantially thicker than it actually is and thus
provides a false sense of security.
Ground water has different origins; however, all fresh
ground water originated from precipitation that
infiltrated. Magmatic or juvenile water is "new" water
that has been released from molten igneous rocks.
The steam that is so commonly given off during
volcanic eruptions is probably not magmatic, but
rather shallow ground water heated by the molten
magma. Connate ground water is defined as that
entrapped within sediments when they were
deposited. Ground water, however, is dynamic and
there is probably no connate water that meets this
definition. Rather, the brines that underlie all or nearly
all fresh ground water have changed substantially
through time because of chemical reactions with the
geologic framework.
4.5.2 Aquifers and Aquitards
In the subsurface, rocks serve either as confining
units or as aquifers. A confining unit or aquitard is
characterized by low permeability that does not
readily permit water to pass through it despite the fact
that it stores large quantities of water. Examples
include shale, clay, and silt. An aquifer has sufficient
permeability to permit water to flow through it with
relative ease and, therefore, it will provide a usable
quantity to a well or spring.
Water occurs in aquifers under two different
conditions - unconfined and confined (Figure 4-
26). An unconfined or water-table aquifer has a free
water surface that rises and falls in response to
differences between recharge and discharge. A
confined or artesian aquifer is overlain by an aquitard
and the water is under sufficient pressure to rise
above the base of the confining bed if it is perforated.
In some cases, the water is under enough pressure
to rise to some point above land surface. This is
called a flowing artesian well. The water level in an
unconfined aquifer is referred to as the water table;
with confined aquifers the water level is called the
potentiometric surface.
Water will arrive at some point in an aquifer through
one of several means. The major source is direct
infiltration of precipitation, which occurs nearly
everywhere. Where the water table lies below a
stream or canal, water will infiltrate. This source is
important part of the year in some places and is a
continuous source in others. Interaquifer leakage, or
flow from one aquifer to another, is probably the most
significant source in deeper, confined aquifers.
Likewise, leakage from aquitards is very important
where pumping from adjacent aquifers has lowered
the head sufficiently for leakage to occur. Underflow,
which is the normal movement of water through an
aquifer, will also transport ground water to a specific
point. Additionally, water can reach an aquifer through
73
-------
Figure 4-26 Aquifer A is unconfined and aquifers B and C
are confined, but water may leak through
confining units to recharge adjacent water-
Table 4-1 Selected Values of Porosity, Specific Yield,
and Specific Retention.
;• Water-o/'-J'
:" Table'-£*•>'
" -
•£-:^r.:-.:- •'"^*-;.
^:.^%^^>^-
/••*:."-'; Aquifer A '-
•* - '*":.".*• •!• »• •'„*••"•
•y.v'l"-'. Aquifer B '-.7
.
V.'t'" Aquifer C
Aquifer A is unconfined and aquifers B and C are
confined, but water may leak through confining units to
recharge adjacent water-bearing zones.
artificial means, such as leakage through ponds, pits,
and lagoons.
An aquifer serves two functions; one as a conduit
through which flow occurs, and the other as a storage
reservoir. This is accomplished by means of openings
in the rock. The openings include those between
individual grains and those present in joints, fractures,
tunnels, and solution openings. There are also
artificial openings, such as engineering works,
abandoned wells, and mines. The openings are
primary if they were formed at the time the rock was
emplaced; they are secondary if they developed after
lithification. Examples of the latter include fractures
and solution openings.
4.5.3 Porosity and Hydraulic Conductivity
Porosity, expressed as a percentage or decimal
fraction, is the ratio between the openings and the
total rock volume. It defines the amount of water a
saturated rock volume can contain. If a unit volume of
saturated rock is allowed to drain by gravity, not all of
the water it contains will be released. The volume
drained is the specific yield, a percentage, and the
volume retained is the specific retention. It is the
specific yield that is available to wells. Therefore,
porosity is equal to specific yield plus specific
retention. Typical values for various rock types are
listed in Table 4-1.
Permeability (P) is used in a qualitative sense, while
hydraulic conductivity (K) is a quantitative term. They
are often expressed in units of gpd/ft2 (gallons per
day per square foot) and refer to the ease with which
water can pass through a rock unit. It is the hydraulic
conductivity that allows an aquifer to serve as a
conduit. Hydraulic conductivity values range widely
from one rock type to another and even within the
Material
Soil
Clay
Sand
Gravel
Limestone
Sandstone, semiconsolidated
Granite
Basalt, young
Porosity
55
50
25
20
20
11
0.1
11
Specific Yield
(% by vol)
40
2
22
19
18
6
0.09
8
Specific
Retention
15
48
3
1
2
5
0.01
3
same rock. Those rocks or aquifers in which the
hydraulic conductivity is nearly uniform are called
homogeneous and those in which it is variable are
heterogeneous or nonhomogeneous. Hydraulic
conductivity can also vary horizontally, in which case
the aquifer is anisotropic. If uniform in all directions,
which is rare, it is isotropic. The fact that both
unconsolidated and consolidated sedimentary strata
are deposited in horizontal units is the reason that
hydraulic conductivity is generally greater horizontally
than vertically by at least an order of magnitude.
Typical ranges in values of hydraulic conductivity for
most common water-bearing rocks are shown in
Table 1-3.
4.5.4 Hydraulic Gradient
The hydraulic gradient (I) is the slope of the water
table or potentiometric surface and is the change in
water level per unit of distance along the direction of
maximum head decrease. It is determined by
measuring the water level in several wells. The water
level in a well, usually expressed as feet above sea
level, is the total head (ht), which consists of
elevation head (z) and pressure head (hp).
ht = z + hp (4-6)
The hydraulic gradient is the driving force that causes
ground water to move in the direction of decreasing
total head. It is generally expressed in consistent
units such as feet per foot. For example, if the
difference in water level in two wells 1,000 feet apart
is 8 feet, the gradient is 8/1,000 or 0.008. The
direction of ground-water movement and the
hydraulic gradient can be determined by information
from three wells (Figure 4-27).
4.5.5 Potentiometric Surface Map
A potentiometric surface or water-level map is a
graphical representation of the gradient. One can be
prepared by plotting water-level measurements on a
base map and then drawing contours. The map
should be drawn so that it actually reflects the
hydrogeological conditions. An example is shown in
Figure 4-28.
74
-------
Figure 4-27 The generalized direction of ground-water
movement can be determined by means of the
water level in three wells of similar depth
(from Heath and Trainer, 1981).
Direction of Ground-
Water Movement
27.6
Water Table Altitude
27.5 . -f / *c
Segments of
Water Table Contours
27.2
27.0 X- I
26.8
The contours are called equipotential lines, indicating
that the water has the potential to rise to that
elevation. In the case of a confined aquifer, however,
the water may have the potential to rise to a certain
elevation, but it cannot actually do so until the
confining unit is perforated by a well. Potentiometric
surface maps are an essential part of any ground-
water investigation because they indicate the direction
in which ground water is moving and provide an
estimate of the gradient, which controls velocity.
A potentiometric surface map can be developed into
a flow net by constructing flow lines that intersect the
equipotential lines or contour lines at right angles.
Flow lines are imaginary paths that would be followed
by particles of water as they flow through the aquifer.
Although there are an infinite number of both
equipotential and flow lines, the former are
constructed with uniform differences in elevation
between them and the latter so that they form, in
combination with equipotential lines, a series of
squares. A carefully prepared flow net in conjunction
with Darcy's Law (discussed below) can be used to
estimate the quantity of water flowing through an
area.
Figure 4-28 A potentiometric surface map representing the hydraulic gradient.
\
\ s~ Aquifer Boundary
>/
\
> 638 Well Location and Altitude of Water Level (feet)
75
-------
4.5.6 Calculating Ground-Water Flow
Darcy's Law, expressed in many different forms, is
used to calculate the quantity of underflow or vertical
leakage. One means of expressing it is:
Q = KIA
(4-7)
where:
Q =
K =
I =
A =
quantity of flow per unit of time, in gpd
hydraulic conductivity, in gpd/ft2
hydraulic gradient, in ft/ft
cross-sectional area through which the flow
occurs, in ft2
The flow rate is directly proportional to the gradient
and therefore the flow is laminar, which means the
water will follow distinct flow lines rather than mix with
other flow lines. Where this does not occur, as in the
case of unusually high velocity which might be found
in fractures, solution openings, or adjacent to some
pumping wells, the flow is turbulent.
Notice in Figure 4-29a that a certain quantity of fluid
(Q) enters the sand-filled tube and the same amount
exits. The water level declines along the length of the
flow path (L) and the head is higher in the manometer
at the beginning of the flow path than it is at the other
end. The difference in head (H) along the flow path
(L) is the hydraulic gradient (H/L or I). The head loss
reflects the energy required to move the fluid this
distance. If Q and A remain constant but K is
increased, then the head loss decreases. It is
particularly important to keep in mind that the head
loss occurs in the direction of flow.
In Figure 4-29b, the flow tube has been inverted and
the water is flowing from bottom to top or top to
bottom. Q, K, A, and I all remain the same. This
illustrates an important concept when the
manometers are considered as wells. Notice that the
deeper well has a head that is higher than the shallow
well when the water is moving upward, whereas the
opposite is the case when the flow is downward.
Where this occurs in the field, it clearly shows the
existence of recharge and discharge areas. In
recharge areas, shallow wells have a higher head
than deeper wells; the difference indicates the energy
required to vertically move the water the distance
between the screens of the two wells. Where the flow
is horizontal, there should be no difference in head.
Along stream valleys, which are regional discharge
areas, the deeper well will have the higher head
(Figure 4-29c). The location of waste disposal sites
in recharge areas might lead to the vertical migration
of leachate to deeper aquifers, and from this
perspective, disposal sites should be locatedin
discharge areas.
An example of the use of Darcy's Law is shown in
Figure 4-30. In this case, a sand aquifer about 30
feet thick lies within the flood plain of a river that is
about a mile wide. The aquifer is covered by a
confining unit of glacial till, the bottom of which is
about 45 feet below land surface. The difference in
water level in two wells a mile apart is 10 feet. The
hydraulic conductivity of the sand is 500 gpd/ft2. The
quantity of underflow passing through cross section
A-A' (Figure 4-30) is:
Q = KIA
= 500 gpd/ft2 x (10 ft/5280 ft) x (5280 X 30)
= 150,000 gpd
Ground water moves both through aquifers and
confining units. Because the difference in hydraulic
conductivity between aquifers and confining units
commonly differs by several orders of magnitude, the
head loss per unit of distance in an aquifer is far less
than in a confining unit. Consequently, lateral flow in
confining units is small compared to aquifers, but
vertical leakage through them can be significant.
Owing to the large differences in hydraulic
conductivity, flow lines in aquifers tend to parallel the
boundaries but in confining units they are much less
dense (Figure 4-31). The flow lines are refracted at
the boundaries in order to produce the shortest flow
path in the confining unit. The angles of refraction are
proportional to the differences in hydraulic
conductivity.
If one is concerned about the flow from one aquifer to
another via a confining unit, a slightly modified form of
Darcy's Law can be used:
QL = (p/m)AH
(4-8)
where:
QL = quantity of leakage, in gpd
p = vertical hydraulic conductivity of the
confining unit, in gpd/ft2
m = thickness of the confining unit, in ft
A = cross sectional area, in ft2
H = difference in head between the two wells
As illustrated in Figure 4-32, assume two aquifers
are separated by a layer of silt. The silty confining
unit is 10 feet thick and has a vertical permeability of
2 gpd/ft2. The difference in water level in wells
tapping the upper and lower aquifers is 15 feet. Let
us also assume that these hydrogeologic conditions
exist in an area of 1 square mile. The daily quantity of
leakage that occurs within this area from the
shallower aquifer to the deeper one is:
QL = (2 gpd/ft2/10 ft) x 52802 x 15 ft
= 83,635,200 gpd
This calculation clearly shows that the quantity of
leakage, either upward or downward, can be highly
76
-------
Figure 4-29 Graphical explanation of Darcy's Law. Notice that the flow in a tube can be horizontal or vertical in the
direction of decreasing head.
A. Horizontal sand-filled tube.
Hf
•Q
Gradient = H/L = I, the energy required
to move the water distance L
Q = Quantity of flow, gpd
A = Cross sectional area of flow, ft2
K = Hydraulic conductivity = gpd/ft2
B. Vertical tube with flow
from bottom to top.
Q
—
L
I
Vertical tube with flow
from top to bottom.
u
L
1
-
C. Field conditions.
Recharge
Area
Water Level
Horizontal
Flow
Discharge
Area
lit
^
. 1.
^ ,
. 1-
• ;
/
/
/
77
-------
Figure 4-30 Using Darcy's Law to estimate underflow in an
aquifer.
Well!
Well 2
Glacial Till
(Clay) 45' 2_
Sand 30'
18'
significant even if the hydraulic conductivity of the
confining unit is small.
4.5.7 Interstitial Velocity
The interstitial velocity of ground water is of particular
importance in contamination studies. It can be
estimated by the following equation:
v = Kl/7.48n
(4-9)
where:
v = average velocity, in ft/d
n = effective porosity
Other terms are as previously defined.
As an example, assume there is a spill that consists
of a conservative substance such as chloride. The
liquid waste infiltrates through the unsaturated zone
and quickly reaches a water-table aquifer that
consists of sand and gravel with hydraulic
Figure 4-31 Long-term ground-water hydrographs show that the water level fluctuates in response to differences
between recharge and discharge.
. Shale — - — —"
Water Table
Equipotential LinesXr -- -/Flow Lines
Head above - f
the Datum _~ J~ ~
Plane
(2)
Bedrock
78
-------
Figure 4-32 Using Darcy's Law to calculate the quantity of
leakage from one aquifer to another.
Aquifer
Aquifer
— —
0 0 0 • 0
0 Q 0 Q 0
*> 0 00
3:
» 0
'•'o<
0 0
z-z-z-i
o , ° ° o o " , °
0 » ' * ° ° 0 0
0 » , . P° » '
o
0
> g
—
0
O 0
b
0 0
.'"'f '•*•
.•.I.- 1 •
~ — _ r~
0 • 00°
« ° « 0 . °
o o e o t> «
Area of leakage = 1 mi2
P1 = 2gpd/ft2
m1 = 10ft
Ah = 15ft
Q = PIA = -^p AAh
m
Q = ~ x (5280 x 5280} x 15 = 83,635,200 gpd
conductivity of 2,000 gpd/ft2 and effective porosity of
0.20. The water level in a well at the spill lies at an
altitude of 1,525 feet and, at a well a mile directly
downgradient, it is at 1,515 feet (Figure 4-33). What
is the velocity of the water and contaminant and how
long will it be before the second well is contaminated
by chloride?
v = (2000 gpd/ft2) x (10 ft/5280 ft)/7.48 x .20
= 2.5 ft/d
Time = 5280 ft/2.5 ft/d
= 2112 days or 5.8 yr
This velocity value is crude at best and can only be
used as an estimate. Hydrodynamic dispersion, for
example, is not considered in the equation. This
phenomenon causes particles of water to spread in a
direction that is transverse to the major direction of
flow and to move downgradient at a rate faster than
expected. It is caused by an intermingling of
streamlines due to differences in interstitial velocity
brought about by the irregular pore space and
interconnections.
Furthermore, most chemical species are retarded in
their movement by reactions with the geologic
framework, particularly with certain clays, soil-
organic matter, and certain hydroxides. Only
conservative substances such as the chloride ion will
move unaffected by retardation.
In addition, it is not only the water below the water
table that is moving, but also fluids within the capillary
Figure 4-33 Using ground-water velocity calculations, it
would require nearly six years for a
contaminant to reach the downgradient well
under the stated conditions.
Gasoline Spill
" Sand and Gravel
P = 2000
oa = 20%
, Table
• 9
V =
PI
10
2000 x 5280
-3JL
1.5
= 2.5 ft/day
Time =
7.48a 7.48 x 0.2
5280'
2.5 ft/day
It would require 5.8 years for gasoline to reach downgradient under
existing conditions.
= 2112 daVs or 5-8 years
fringe. Here the velocity diminishes rapidly upward
from the water table. Movement in the capillary fringe
is important where the contaminant is gasoline or
other substances less dense than water.
4.5.8 Transmissivity and Storativity
Hydrogeologists commonly use the term
transmissivity (T) to describe the capacity of an
aquifer to transmit water. Transmissivity is equal to
the product of the aquifer thickness (m) and hydraulic
conductivity (K) and is measured in units of gpd/ft of
aquifer thickness:
T = Km
(4-11)
Another important term is Storativity (S), which
describes the quantity of water that an aquifer will
release from or take into storage per unit surface area
of the aquifer per unit change in head. In unconfined
aquifers the storagtivity is, for all practical purposes,
equal to the specific yield and, therefore, should
range between 0.1 and 0.3. The Storativity of confined
aquifers is substantially smaller because the water
which is released from storage when the head
declines comes from the expansion of water and
compression of the aquifer, both of which are very
small. For confined aquifers, Storativity generally
ranges between 0.0001 and 0.00001; for leaky
confined aquifers it is in the range of 0.001. The small
Storativity for confined aquifers means that to obtain a
sufficient supply from a well there must be a large
pressure change throughout a wide area. This is not
the case with unconfined aquifers because the water
derived is not related to expansion and compression
but comes instead from gravity drainage and
dewatering of the aquifer.
Hydrogeologists have found it necessary to use
transmissivity and Storativity to calculate the response
79
-------
of an aquifer to stresses and to predict future water
level trends. These terms are also required as input
for most flow and transport computer models.
4.5.9 Water-Level Fluctuations
Ground-water levels fluctuate throughout the year in
response to natural changes in recharge and
discharge (or storage), to changes in pressure, and to
artificial stresses. Fluctuations brought about by
changes in pressure are limited to confined aquifers.
Most of these changes are short term and are caused
by loading, such as a passing train compressing the
aquifer, or by an increase in discharge from an
overlying stream. Others are related to changes in
barometric pressure, tides, earthtides, and
earthquakes. None of these fluctuations reflect a
change in the volume of water in storage.
Fluctuations that involve changes in storage are
generally more long lived (Figure 4-34). Most
ground-water recharge takes place during the
spring, which causes the water level to rise. Following
this period, which is a month or two long, the water
level declines in response to natural discharge, which
is largely to streams. Although the major period of
recharge occurs in the spring, minor events can
happen any time there is a rain.
Evapotranspiration effects on a surficial or shallow
aquifer are both seasonal and daily. Plants, each
serving as a minute pump, remove water from the
capillary fringe or even from beneath the water table
during hours of daylight in the growing season. This
results in a diurnal fluctuation in the water table and
stream flow.
4.5.70 Cone of Depression
When a well is pumped, the water level in its vicinity
declines to provide a gradient to drive water toward
the discharge point. The gradient becomes steeper as
the well is approached because the flow is
converging from all directions and the area through
which the flow is occurring gets smaller. This results
in a cone of depression around the well (Figure 4-
35). Relatively speaking, the cone of depression
around a well tapping an unconfined aquifer is small if
compared to that around a well in a confined system.
The former may be a few tens to a few hundred feet
in diameter, while the latter may extend outward for
miles.
Cones of depression from several pumping wells may
overlap and, since their drawdown effects are
additive, the water-level decline throughout the area
of influence is greater than from a single cone (Figure
4-36). In ground-water studies and particularly
contamination problems, evaluation of the cone or
cones of depression can be critical because they
represent an increase in the hydraulic gradient, which
in turn controls ground-water velocity and direction
of flow. In fact, properly spaced and pumped wells
provide a mechanism to control the migration of
leachate plumes. Discharging and recharging well
schemes are commonly used in attempts to restore
contaminated aquifers.
4.5.11 Specific Capacity
The decline of the water level in a pumping well, or
any well for that matter, is called the drawdown and
the prepumping level is the static water level.
(Figure 4-37). The discharge rate of the well divided
by the difference between the static and the pumping
level is the specific capacity. The specific capacity
indicates how much water the well will produce per
foot of drawdown:
Specific capacity = Q/s
(4-11)
where:
Q = discharge rate, in gpm
s = drawdown, in ft
If a well produces 100 gpm and the drawdown is 8 ft,
the well will produce 12.5 gpm for each foot of
available drawdown. One can rather crudely estimate
transmissivity by multiplying specific capacity by
2,000.
80
-------
Figure 4-34 Long-term ground-water hydrographs show that the water level fluctuates in response to differences
between recharge and discharge.
30
OJ
| 40
3
W
•D 50
JO
3 60
70
•S 80
2 90
(0
f 110
IB
Q
120
In San Antonio, Bexar County
} '"' Edwards and Associated Limestones
1956 57 58 59 60 61 62 63 64 65 66 67 68 69 70 71 72 73 74 75 1976
Years
Figure 4-35 Cones of depression in unconfined and confined aquifers (from Heath, 1983).
Land Surface
Limits of Cone
of Depression
Land Surface -
Potentiometric Surface-
Q ~"~
// s"
/ / s
//'
Drawdown
Confining Bed
////////////
1 ^^ N.\
T
\
x
/ / //
o ^
O -ff
Confined Aquifer
o — ^f
i
P
r.t
-------
Figure 4-36 Overlapping cones of depression result in more drawdown than would be the case for a single well (from
Heath, 1983).
Well
A
Well
B
Static Potentiometric Surface
Cone of
Depression with •
Well A Pumping
Cone of Depression if
Well B Were Pumping
and Well A Were Idle
////////////S///////S//S1
Confined Aquifer
Well
A
Well
B
////////// / / / /
Cone of Depression
• with Both Well A and
B Pumping
Confined Aquifer
//////////'
4.6 References
Cross, W.P., and R.E. Hedges. 1959. Flow Duration
of Ohio Streams. Ohio Division of Water Bulletin 31.
Daniel, J.F., L.W. Cable, and RJ. Wolf. 1970. Ground
Water - Surface Water Relation During Periods of
Overland Flow. U.S. Geological Survey Professional
Paper 700-B, U.S. Government Printing Office,
Washington, D.C.
Durfor, C.N., and E. Becker. 1962. Public Water
Supplies of the 100 Largest Cities in the United
States. U.S. Geological Survey Water-Supply Paper
1812. U.S. Government Printing Office, Washington,
D.C.
Freeze , R.A., and J.A. Cherry. 1979. Groundwater.
Prentice-Hall Publishing Co., Inc., Englewood Cliffs,
NJ.
Heath, R.C. 1984. Ground-Water Regions of the
United States. U.S. Geological Survey Water-Supply
Paper 2242, U.S. Government Printing Office,
Washington, D.C.
Heath, R.C. 1983. Basic Ground-Water Hydrology.
U.S. Geological Survey Water-Supply Paper 2220,
U.S. Government Printing Office, Washington, D.C.
Heath, R.C., and F.W. Trainer. 1981. Introduction to
Ground Water Hydrology. Water Well Journal
Publishing Co., Worthington, OH.
Johnson, E.E. 1966. Ground-Water and Wells.
Edward E. Johnson, Inc., Saint Paul, MN.
Kunkle, G.R. 1965. Computation of Ground-Water
Discharge to Streams During Floods, or to Individual
Reaches During Base Flow, by Use of Specific
Conductance. U.S. Geological Survey Professional
Paper 525-D, U.S. Government Printing Office,
Washington, D.C.
LaSala, A.M. 1967. New Approaches to Water-
Resources Investigations in Upstate New York.
Ground Water 5(4).
Meinzer, O.E., and N.D. Stearns. 1928. A Study of
Ground Water in the Pomerang Basin. U.S.
82
-------
Figure 4-37 Values of transmissivrty based on specific capacity are commonly too small because of well construction
details (from Heath, 1983).
Land Surface
Potentiometric Surface
Cone of Depression
Confining Bed
(Nonpumping)
Drawdown in
the Aquifer
B. Magnitude of the Well Loss Compared to the
Drawdown in the Aquifer
Well
Loss
A. Thickness of the Producing Zone Compared to the
Length of the Screen or Open Hole
C. The Difference Between the "Nominal" Radius and
the Effective Radius
"Nominal"
Radius
Producing
Zone
Length of
Screen
o
o
o
o
o
o
o
i. °
o
Effective
Radius
Confined
Aquifer
Geological Survey Water-Supply Paper 597-B,
U.S. Government Printing Office, Washington, D.C.
Olmsted, F.H., and A.G. Hely. 1962. Relation
Between Ground Water and Surface Water in
Brandywine Creek Basin, Pennsylvania. U.S.
Geological Survey Professional Paper 417-A, U.S.
Government Printing Office, Washington, D.C.
Pettyjohn, W.A. 1985. Regional Approach to
Ground-Water Investigations. In: Ground Water
Quality, edited by C.H. Ward, W. Giger, and P.L.
McCarty, John Wiley & Sons, New York, NY
Pettyjohn, W.A., H. White, and S. Dunn. 1983. Water
Atlas of Oklahoma. University Center for Water
Research, Oklahoma State University, OK
Pettyjohn, W.A. 1982. Cause and Effect of Cyclic
Changes in Ground-Water Quality. Ground-Water
Monitoring Review 2(1).
Pettyjohn, W.A. and R.J. Henning. 1979. Preliminary
Estimate of Ground-Water Recharge Rates, Related
Streamflow and Water Quality in Ohio. Project
Completion Report 552, Ohio State University Water
Resources Center, OH.
Pettyjohn, W.A. 1975. Chloride Contamination in Alum
Creek, Central Ohio. Ground Water 13(4).
Rasmussen, W.C., and G.E. Andreason. 1959.
Hydrologic Budget of the Beaver Dam Creek Basin,
Maryland. U.S. Geological Survey Water-Supply
Paper 1472, U.S. Government Printing Office,
Washington, D.C.
Seaber, P.R. 1965. Variations in Chemical Character
of Water in the Englishtown Formation of New Jersey.
U.S. Geological Survey Professional Paper 498-B,
U.S. Government Printing Office, Washington, D.C.
83
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Schicht, R.J., and W.C. Walton. 1961. Hydrologic
Budgets for Three Small Watersheds in Illinois. Illinois
State Water Survey Report of Investigation 40.
Stefferud, Alfred. 1955. Water, the Yearbook of
Agriculture. U.S. Department of Agriculture.
Todd, O.K. 1980. Groundwater Hydrology. John Wiley
& Sons, New York, NY.
Toler, L.G. 1965. Use of Specific Conductance to
Distinguish Two Base-Flow Components in Econfina
Creek, Florida. U.S. Geological Survey Professional
Paper 525-C, U.S. Government Printing Office,
Washington, D.C.
Trainer, F.W., and F.A. Watkins. 1975. Geohydrologic
Reconnaissance of the Upper Potomac River Basin.
U.S. Geological Survey Water-Supply Paper 2035,
U.S. Government Printing Office, Washington, D.C.
U.S. Environmental Protection Agency. 1985.
Protection of Public Water Supplies from Ground-
Water Contamination. EPA-625/4-85-016, Center
for Environmental Research Information, Cincinnati,
OH.
U.S. Geological Survey. 1985. Water Resources
Data, Oklahoma, Water Year 1983. U.S. Geological
Survey Water-Data Report OK-83-1, U.S.
Government Printing Office, Washington, D.C.
Viscoky, A.P. 1970. Estimating the Ground-Water
Contribution to Storm Runoff by Electrical
Conductance Method. Ground Water 8(2).
84
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CHAPTERS
MONITORING WELL DESIGN AND CONSTRUCTION
The principal objective of constructing monitoring
wells is to provide access to an otherwise
inaccessible environment. Monitoring wells are used
to evaluate topics within various disciplines, including
geology, hydrology, chemistry, and biology. In
ground-water quality monitoring, wells are used for
collecting ground-water samples, which upon
analysis, may allow description of a contaminant
plume, or the movement of a particular chemical (or
biological) constituent, or ensure that potential
contaminants are not moving past a particular point.
5.1 Ground-Water Monitoring Program
Goals
Each purpose for ground-water monitoring, ambient
monitoring, source monitoring, case preparation
monitoring, and research monitoring, (Barcelona et
a/., 1983) must satisfy somewhat different
requirements, and may require different strategies for
well design and construction. At the outset, it must be
clearly understood what the intended monitoring
program is to accomplish and the potential future use
of the wells in other, possibly different, monitoring
programs.
Regional investigations of ground-water quality fall
into the ambient monitoring category. Such
investigations seek to establish an overall picture of
the quality of water within all or portions of an aquifer.
Generally, sample collection is conducted routinely
over a period of many years to determine changes in
quality over time. Often, changes in quality are related
to long-term changes in land use (e.g., the effects of
urbanization). Monitoring conducted for Safe Drinking
Water Act compliance generally falls in this category.
Samples are often collected from a variety of public
and private water supply wells for ambient quality
investigations. Because of this, the data obtained
through some ambient monitoring programs may not
meet the strict well design and construction
requirements imposed by the three other types of
monitoring. However, such programs are important for
detecting significant changes in aquifer water quality
over time and space and protecting public health.
Regulatory monitoring at potential contaminant
sources is considered source monitoring. Under this
type of program, monitoring wells are located and
designed to detect the movement of specific
pollutants outside the boundaries of a particular
facility (e.g., treatment, storage, or disposal).
Ground-water sampling to define contaminant plume
extent and geometry would fall into this classification
of monitoring. Monitoring well design and construction
are tailored to the site geology and contaminant
chemistry. Quantitative aspects of analytical results
become most important because the level of
contaminant concentration may require specific
regulatory action.
Monitoring for case preparation, such as for legal
proceedings in environmental enforcement, requires a
level of detail similar to source monitoring. Source
monitoring, in fact, often becomes part of legal
proceedings to establish whether or not
environmental damage has occurred and identify the
responsible party. This is a prime example of one
type of monitoring program evolving into another. The
appropriateness and integrity of monitoring well
design and construction methods will come under
much scrutiny. In such cases, the course of action
taken during the monitoring investigation, the
decisions that were made concerning well design and
construction, and the reasons why those decisions
were made must be clearly established and
documented.
Monitoring for research generally requires a level of
sophistication beyond that required of any other type
of monitoring (this, of course, depends upon the
types and concentrations of constituents being sought
and the overall objectives of the research). Detailed
information is often needed to support the basic
concepts and expand the levels of understanding of
the complex mechanisms of ground-water
movement and solute/contaminant transport.
The goals of any proposed ground-water monitoring
program should be clearly stated and understood
before decisions are made on the types and numbers
of wells needed, where they should be located, how
deep they should be, what constituents are of
interest, and how samples should be collected,
stored, transported, and analyzed.
85
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As each of these decisions is made, consideration
must be given to the costs involved in each step of
the monitoring program and how compromises in one
step may affect the integrity and outcome of the other
steps. For example, cost savings in well construction
materials may so severely limit the usefulness of a
well that another well may need to be constructed at
the same location for the reliable addition of a single
chemical parameter.
5.2 Monitoring Well Design Components
Monitoring well design and construction methods
follow production well design and construction
techniques; however, it must be remembered that a
monitoring well is built specifically to give access to
the ground water so a "representative" sample of
water can be withdrawn and analyzed. While it is
important to pay attention to well efficiency and yield,
the ability to produce large amounts of water for
supply purposes is not the primary objective.
Emphasis, then, is placed on constructing a well that
will provide easily obtainable ground-water samples
that will give reliable, meaningful information. It
follows from this emphasis that the materials and
techniques used for constructing a monitoring well
must not materially alter the quality of the water being
sampled. An understanding of the chemistry of
suspected pollutants and the geologic setting in which
the monitoring well is to be constructed play a major
role in the drilling technique and well construction
materials used.
There are several components to be considered in
monitoring well design. These include: location (and
number of wells), diameter, casing and screen
material, screen length and depth of placement,
sealing material, well development, and well security.
Often, discussion of one component will impinge
upon other components.
5.2.7 Location and Number
Locating monitoring wells spatially and vertically to
ensure that the ground- water flow regime of
concern is being monitored is obviously one of the
most important components in ground-water quality
monitoring design. It is impossible to divorce
prescribing monitoring well locations (sites) and the
number of wells in the monitoring program. The
number of wells and their location are principally
determined by the purpose of the monitoring program.
In most monitoring situations, the goal is to determine
the effect some surface or near-surface activity has
had on nearby ground-water quality. Most dissolved
constituents will descend vertically through the
unsaturated zone beneath the area of activity and
then, upon reaching the saturated zone, move
horizontally in the direction of ground-water flow.
Therefore, monitoring wells are normally completed
downgradient in the first permeable water-bearing
unit encountered. Consideration should be given to
natural (seasonal) and artificial fluctuations in water
table elevation. Artificial fluctuations include pumpage,
which will cause water levels to fall, and lagoon
operation, which can cause a rise or "mound" in the
water table.
Preliminary boreholes and/or monitoring wells can be
constructed for the collection and analysis of geologic
material samples, ground-water levels, and water
quality samples to guide the placement of additional
wells. Accurate water level information must be
established to determine if local ground-water flow
paths and gradients differ significantly from the
regional appraisal.
The analysis of water quality samples from the
preliminary wells can also direct the placement of
additional wells. Such wells are particularly helpful in
the vertical arrangement of sampling points
(especially for a contaminant that is denser than
water). Without some preliminary chemical data, it is
usually very difficult to know where the most
contaminated zone is.
A number of factors will govern where and how many
wells should be constructed. These factors include:
site geology, site hydrology, source characteristics,
contaminant characteristics, and the size of the area
under investigation. Certainly, the more complicated
the geology and hydrology, the more complex the
contaminant and source, and the larger the area
being investigated, the greater the number of
monitoring wells that will be required. Details of some
of these factors are discussed in Chapters 1, 2, 4,
and 9 and in the following sections.
5.2.2 Diameter
In the past, the diameter of a monitoring well was
based primarily on the size of the device (bailer,
pump, etc.) being used to withdraw the water
samples. This practice was similar to that followed for
water supply well design. For example, a domestic
water well is commonly 4 to 6 inches in diameter to
accommodate a submersible pump capable of
delivering from 5 to 20 gallons per minute. Municipal,
industrial, and irrigation wells have greater diameters
to handle larger pumps and to increase the available
screen open area so the well can produce water
efficiently.
This practice worked well in very permeable
formations, where an aquifer capable of furnishing
large volumes of water was present. However, unlike
most water supply wells, monitoring wells are quite
often completed in very marginal water-producing
zones. Pumping one or more well volumes of water
(the amount of water stored in the well casing under
nonpumping conditions) from a well built in low-
yielding materials (Gibb et a/., 1981) may present a
serious problem if the well has a large diameter.
86
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Figure 5-1 illustrates the amount of water in storage
per foot of casing for different well casing diameters.
Well casings with diameters of 2 to 6 inches will
contain 0.16 and 1.47 gallons of water per foot of
casing, respectively. Purging four well volumes from a
well containing 10 feet of water would require removal
of 6.4 gallons of water from a 2-inch well and 58.8
gallons of water from a 6-inch well. Under low-
yielding conditions, it can take considerable lengths of
time to recover enough water in the well to collect a
sample (Figure 5-2).
In addition, when hazardous constituents are present
in the ground water, proper disposal of the purged
water will be necessary. This amount of water should
be kept to a minimum, for safety's sake as well as
disposal cost. Cost of well construction is also a
consideration. Small diameter wells (less than 4
inches) are much less expensive than large diameter
wells in terms of both cost of materials and cost of
drilling.
For these reasons and with the advent of numerous
commercially available small-diameter pumps (less
than 2 inches OD) capable of lifting water over 100
feet, 2-inch ID monitoring wells have become the
standard in monitoring well technology.
Large diameter wells can be useful in situations
where monitoring may be followed by remedial
actions involving reclamation and treatment of the
contaminated ground water. In some instances, the
"monitoring" well may become a "supply" well to
remove contaminated water from the ground for
treatment. Larger diameter wells also merit
consideration when monitoring is required at depths
of hundreds of feet and in other situations where the
additional strength of large diameter casing is needed.
For sampling at several depths beneath one location,
several monitoring wells have been nested in a single
borehole (Johnson, 1983). A technique such as this
will require drilling a larger diameter hole to
accommodate the multiple well casings. Again, the
use of smaller diameter casing provides advantages
by allowing more wells to be nested in the borehole,
thus easing construction and saving costs in drilling
expenses.
5.2.3 Casing and Screen Material
The type of material used for a monitoring well can
have a distinct effect on the quality of the water
sample to be collected (Barcelona ef a/., 1983;
Gillham et a/., 1983 and Miller, 1982). The materials
of choice should retain their structural integrity for the
duration of the monitoring program under actual
subsurface conditions. They should neither adsorb
nor leach chemical constituents which would bias the
representativeness of the samples collected.
Galvanized steel casing can impart iron, manganese,
zinc, and cadmium to many waters. Steel casing may
impart iron and manganese to a sample. PVC pipe
Figure 5-1 Volume of water stored per foot of well
casing for different diameter casings (from
Rinaldo-Lee, 1983).
2.5
.S 2.0
CO
CO
O
o
1.0
0.5
Figure 5-2
12345678
Well Diameter (Inches)
Time required for recovery after slug of water
removed (from Rinaldo-Lee, 1983).
40
10
1234
Well Diameter (Inches)
Assumptions: K= 1 x 10'5 cm/sec, well screen = 10', 10' of water
above screen, 6' of water instantaneously
removed
87
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has been shown to release and adsorb trace amounts
of various organic constituents to water after
prolonged exposure (Miller, 1982). PVC solvent
cements used to attach sections of PVC pipe have
also been shown to release significant quantities of
organic compounds.
TeflonR and glass are among the most inert materials
that have been considered for monitoring well
construction. Glass, however, is difficult and
expensive to use under most field conditions. TeflonR
is also very expensive; with technological advances,
TeflonR-coated casings and screens may become
available. Stainless steel also offers desirable
properties from a monitoring perspective, but it too is
expensive.
A reasoned strategy for ground-water monitoring
must consider the effects of contaminated water on
well construction materials. Unfortunately, there is
limited published information on the performance of
specific materials in varied hydrogeologic settings
(Pettyjohn et a/., 1981). A preliminary ranking of
commonly used materials exposed to different
solutions representing the principal soluble species
present in hazardous waste site investigations
produced the following list, in order of best to worst
(Barcelona et a/., 1983):
TeflonR
Stainless Steel 316
Stainless Steel 304
PVC Type I
Lo-Carbon Steel
Galvanized Steel
Carbon Steel
Polyvinyl chloride (PVC Type I) has very good
chemical resistance except to low molecular weight
ketones, aldehydes, and chlorinated solvents. As the
organic content of a solution increases, direct attack
on the polymer matrix or solvent absorption,
adsorption, or leaching, may occur. The only
exception to this observation is Teflon. Provided that
sound construction practices are followed, TeflonR
can be expected to out perform all other casing and
sampling materials (Barcelona et a/., 1983).
Stainless steels are the most chemically resistant of
the ferrous materials. Stainless steel may be sensitive
to the chloride ion, which can cause pitting corrosion,
especially over long term exposures under acidic
conditions. Given the similarity in price, workability,
and performance, the remaining ferrous materials
(lo-carbon, galvanized steel, and carbon) provide
little advantage over one another for casing/screen
construction.
Significant levels of organic components found in
PVC primers and adhesives (such as tetrahydrofuran,
methylethylketone, cyclohexanone, and
methylisobutylketone) were detected in well water
several months after installation (Sosebee, ef a/.,
1982). The presence of compounds such as these
can mask the presence of other similar volatile
compounds (Miller, 1982). Therefore, when using
PVC and other similar materials (e.g., ABS,
polypropylene, or polyethylene) for well construction,
threaded joints are the preferred means for
connecting sections together.
In many situations, it may be possible to compromise
accuracy or precision for initial cost, depending on
the objectives of the monitoring program. For
example, if the contaminants of interest are already
defined and they do not include substances which
might bleed or sorb, it may be reasonable to use
wells cased with a less expensive material.
Wells constructed of less than optimum materials
might be used for sampling if identically constructed
wells are constructed in uncontaminated parts of the
monitored aquifer to provide ground-water samples
for use as "blanks" (Pettyjohn et a/., 1981). However,
such blanks may not adequately address problems of
adsorption on or leaching from the casing material
induced by contaminants in the ground water. It may
be feasible to use two or more kinds of casing
materials in the saturated zone and above the
seasonal high water tables, such as TeflonR or
stainless steel, and use a more appropriate material,
such as PVC or galvanized steel casing, below static
water level.
It must be remembered, however, that trying to save
money by compromising on material quality or
suitability may eventually increase program cost by
causing reanalysis, or worse, monitoring well
reconstruction. Careful consideration is required in
each case, and the analytical laboratory should be
fully aware of the construction materials used.
Care must also be given to the preparation of the
casing and well screen materials prior to installation.
At a minimum, materials should be washed with
detergent and rinsed thoroughly with clean water.
Steam-cleaning and high pressure, hot water
cleaners provide excellent cleaning of cutting oils and
lubricants left on casings and screens after
manufacture (this is particularly true for metal casing
and screen materials). To ensure that these and other
sampling materials are protected from contamination
prior to placement down-hole, materials should be
covered (with plastic sheeting or other material) and
kept off the ground.
All wells should allow free entry of water. The water
produced should be as clear and silt-free as
possible. For drinking water supplies, sediment in the
raw water can create additional pumping and
treatment costs and lead to the general unpalatability
of the water. With monitoring wells sediment-laden
water can greatly lengthen filtering time and create
chemical interference in sample analyses.
88
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Commercially manufactured well screens are
preferred for monitoring well design given that the
proper screen slot size is chosen. Sawed or torch-
cut casing may be appropriate in deposits where
medium to coarse sand or gravel predominates. In
formations where fine sand, silt, and clay
predominate, sawed or torch-cut slots will not be
small (or uniform) enough to retain the materials and
the well may clog. The practice of sawing slots in
PVC pipe should be avoided in monitoring situations
where organic chemicals are of concern because this
procedure exposes fresh surfaces of PVC, increasing
the possibility of releasing compound ingredients or
reaction products.
It may be helpful to have several slot-sized well
screens on-site so the correct manufactured screen
can be placed in the hole after the materials within
the zone of interest have been inspected. Gravel
pack compatible with the selected screen slot size will
further help retain the finer fractions of material and
allow freer entry of water into the well by creating a
zone of higher permeability around the well screen.
For natural-packed wells (no gravel pack), where
relatively homogeneous, coarse materials
predominate, a slot size should be selected that will
retain from 40 to 50 percent of the screened material.
In cases where adequate well development
procedures may be difficult to follow, a screen that
will retain about 70 percent of the screened formation
should be selected. If an artificial pack is used, a
uniform gravel-pack size that is from three to five
times the 50 percent size of the formation and a
screen size that will retain at least 90 percent of the
pack material should be selected (Walker, 1974). The
gravel-pack should be composed of clean, uniform
quartz sand.
Placement of the gravel-pack should be done
carefully to avoid bridging in the hole and to allow
uniform settling around the screen. A tremie pipe can
be used to guide the sand to the bottom of the hole
and around the screen. The pipe should be slowly
lifted as the annulus between the screen and
borehole as the borehole fills. If the depth of water
standing in the annulus is not great, the sand can be
simply poured from the surface. Calculations should
be made to determine what volume of sand will be
required to fill the annulus to the desired depth
(usually about one foot above the top of the screen).
Field measurements should be taken to confirm the
pack has reached this level before backfilling or
sealing procedures start.
5.2.4 Screen Length and Depth of Placement
The length of screen chosen and the depth at which
it is placed in monitoring well design are dependent
on, to a large degree, the behavior of the contaminant
as it moves through the unsaturated and saturated
zones and, again, the goal of the monitoring program.
When monitoring a potable water supply aquifer, the
entire thickness of the water-bearing formation could
be screened (just as a production well would be). For
regional aquifer studies, production wells are
commonly used for sampling. Such samples would
provide water integrated over the depth of the water-
bearing zone(s) and would provide a sample similar in
quality to what would be found in a drinking water
supply.
When sampling specific depth intervals at one
location is necessary, vertical nesting of wells is
common. This technique is often necessary when the
saturated zone is too thick to adequately monitor with
one long screened section (causing dilution of the
collected sample). Contaminants tend to stratify within
the saturated zone; collection of a sample integrated
over a thick zone will give little information on the
depth and concentration that a contaminant may have
reached.
Screen lengths of one to two feet are common in
detailed plume geometry investigations. Thick
aquifers would require that several wells be
completed at different depth intervals. In such
situations (and depending on the magnitude of the
aquifer saturated thickness), screen lengths of no
more than 5 to 10 feet are used. Monitoring wells can
be constructed in separate holes placed closely
together or in one larger diameter hole, as in Figure
5-3. Prevention of the vertical movement of
contaminants in the well bore before and after well
completion may be difficult to achieve since multiple
wells in one hole are difficult to seal. Thus, the drilling
of multiple holes may be required to insure well
integrity. Specially constructed installations have been
developed to sample a large number of points
vertically over short intervals (Morrison, 1981;
Pickens, 1981; and Torstensson, 1984; Figures 5-4
and 5-5).
In other situations, only the first water-bearing zone
encountered will require monitoring (for example,
when monitoring near a potential contaminant source
in a relatively impermeable glacial till). Here, the
"aquifer" or zone of interest may be only 6 inches to
a few feet thick. Screen length should be limited to 1
to 2 feet in these cases to minimize siltation problems
from surrounding fine-grained materials and possible
dilution effects from water contributed by
uncontaminated zones.
Because of the chemical reactions which occur when
ground water contacts the atmosphere, particularly for
volatile compounds, aeration of the screens4 section
should be avoided. Well depth should assure that the
screened section is always fully submerged.
Fluctuations in the elevation of the top of the
saturated zone caused by seasonal variations or
man-induced changes must be considered.
89
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Figure 5-3 Typical multiwell installations (from Johnson, 1983).
Well Nest
Single Borehole
7
7
/
/
3\\ W^
Backfill (Typ.)
Grout Seal (Typ.) . fcr
- Filter Sand (Typ.
Screened Interval
Monitoring for contaminants with densities different
than water calls for special attention. In particular, low
density organic compounds such as gasoline will float
on the ground-water surface (Gillham et a/., 1983).
Monitoring wells constructed for floating contaminants
should contain screens which extend above the zone
of saturation so that these lighter substances can
enter the well. The screen length and position must
accommodate the magnitude and depth of variations
in water table elevation. However, the thickness of
floating products in the well does not necessarily
indicate the thickness of the product in the aquifer.
5.2.5 Sealing Materials and Procedures
It is critical that the screened portion of each
monitoring well access ground water from a specific
depth interval. Vertical movement of ground water in
the vicinity of the well can greatly influence sample
quality (Keith et a/., 1982). Rainwater can infiltrate
backfill, potentially diluting or contaminating samples;
vertical seepage of feachate along the well casing will
also produce unrepresentative samples (this is
particularly important in multilevel installations such as
in Figures 5-3, 5-4, and 5-5). Even more
importantly, the creation of a conduit in the annulus of
the monitoring well that could contribute to or hasten
the spread of contamination is to be strictly avoided.
Several methods have been employeed successfully
to isolate contaminated zones during the drilling
process (Burkland and Raber, 1983; Perry and Hart,
1985).
Monitoring wells are usually sealed with neat cement
grout, dry benonite (powdered, granulated, and
pelletized), or bentonite slurry. Well seals usually
occur at two places within the annulus created by the
drilling operation. One area is within or near the
saturated zone to isolate the screened interval for
sampling. The other is at the ground surface to inhibit
downward leakage of surface contaminants.
The use of bentonite traditionally has been
considered to provide a much better seal than
cement. However, recent investigations on the use of
clay liners for hazardous waste disposal have shown
that some organic compounds migrate through
bentonite with little or no attenuation (K.W. Brown, et
90
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Figure 5-4 Schematic diagram of a multilevel sampling
device (from Pickens, 1981).
Field Installation
r— End Cap
Cross Section of
Sampling Point
Ground Surface '
Male ft Female
V Couplings
- PVC Pipe
- Coupling
> Sampling
P°'n"
PVC Pipo
End Cap
a/., 1983). Therefore, cement may offer some
benefits over bentonite.
Bentonite is most often used as a down-hole seal to
prevent vertical migration within the well annulus.
When bentonite must be placed below the water table
(or where water has risen in the bore hole), it is
recommended that a bentonite slurry be tremied
down the annulus to fill the hole from the bottom
upward. In collapsible material conditions, where the
borehole has collapsed to a point just above the water
table, dry bentonite (granulated or pelletized works
best) can be poured down the hole.
Bentonite clay has appreciable ion exchange capacity
which may interfere with the chemistry of collected
samples when the seal is proximate to the screen or
well intake. Cement grout has been known to
seriously affect the pH of sampled water when
improperly placed. Therefore, special attention and
care should be exercised during placement of a
down-hole seal. Approximately one foot (at a
minimum) of gravel pack or naturally collapsed
material should extend above the top of the well
intake to ensure that the sealing materials do not
migrate downward into the well screen. If the sealing
material is too watery before it is placed down the
hole, settling or migration of sealing materials into the
gravel-pack or screened area may occur and the
fine materials in the seal may penetrate the natural or
artificial pack.
Figure 5-5 Single (a) and multiple (b) installation configurations for an air-lift sampler (from Morrison, 1981).
Gas Entry/ Collection Tubing
Bentonite Surface f—-^^, PVC Casing
faeal--^^
. i Air-Lift Sampler i- » ;
]^. Ground Level ^
1, ' : '.'• '. •' :• • • • '
'.'...'•' ^ V Ground-Water Level • ' • •
i •. '•"•".•'.'• ''.•'. .'••.•' .'..'.;•• '•-/ '•'.• . •'
t '•.••.. .••:•.;•'•. ••.'••;. ' • ;
r ~f~ . ' • .Sand Backfill • • -- '. • • '. •'.•'; ''./.'•'. ''.' :
3 '' ' -' '' •'•''•'.-.•'. '.' Air-Lift Sampler
1 .'.-.'. .'.'•'• . ' • , '• • . \ • •' •
i\- ,
n
v
n
Gas Entry/
,X* Collection Tubes
^f , Bentonite Surface
p>f Seal
' " ". • • . . '
i : • •''"•' • ' •
i . •;•-._ ._ •
| Bentonite Seal
-«_Sand Backfill ' . •
-^ Original Boring
•'
-« — Bentonite Seal
j — —Sand Backfill ' ' '
entonite Seal . '• .
Sand Backfill ,
A. Single Air-Lift Sampler with PVC Casing B. Multiple Air-Lift Samplers in Single Boring
91
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While a neat cement (sand and cement, no gravel)
grout is often recommended, especially for surface
sealing, shrinkage and cracking of the cement upon
curing and weathering can create an improper seal.
Shrink-resistant cement (such as Type K Expansive
Cement) and mixtures of small amounts of bentonite
with neat cement have been used successfully to
help prevent cracking.
5.2.6 Development
Development is a facet of monitoring well
construction that often is overlooked. During the
drilling process, fine-grained materials smear on the
sides of the borehole, forming a mud "cake" that
reduces the hydraulic conductivity of the materials
opposite the screened portion of the well. To facilitate
entry of water into the monitoring well (a particularly
important factor for low-yielding geologic materials),
this mud cake must be broken down and the fine-
grained materials removed from the well or well bore.
Development also removes fluids, primarily water,
which are introduced to the water-bearing formations
during the drilling process.
Additionally, monitoring wells must be developed to
provide water free of suspended solids for sampling.
When sampling for metal ions and other inorganic
constituents, water samples must be filtered and
preserved at the well site at the time of sample
collection. Improperly developed monitoring wells will
produce samples containing suspended sediments
that will both bias the chemical analysis of the
collected samples and frequently cause clogging of
the field filtering mechanisms.
The time and money spent for this important
procedure will expedite sample filtration and result in
samples more representative of water contained in
the formation being monitored. The time saved in field
filtration alone will more than offset the cost of
development.
Successful development methods include bailing,
surging, and flushing with air or water. The basic
principle behind each method is to create reversals of
flow in and out of the well (and/or bore hole) to break
down the mud cake and draw the finer materials into
the hole for removal. This process also helps remove
the finer fraction of materials in proximity to the
borehole, leaving behind a "natural" pack of
coarser-grained materials.
Years ago, small-diameter well development was
most commonly achieved through use of a bailer. The
bailer was about the only "instrument" which had
been developed for use in such wells. Rapidly
dropping and retrieving the bailer in and out of the
water caused a back-and-forth action of water in
the well, moving some of the more loosely bound
fine-grained materials into the well where they could
be removed.
Depending on the depth of water in the well, the
length of the well screen, and the volume of water the
bailer could displace, this method was not always
very efficient. "Surge blocks" which could fit inside
2-inch wells provided some improvement on bailing
techniques. Such devices are simply plungers which,
when given a vigorous up-and-down motion,
transfer that energy to an in-and-out action on the
water near the well screen. Surge blocks have the
potential to move larger quantities of water with
higher velocities but pose some risk to the well
casing and screen if too tight a fit is made or if the
up-and-down action becomes too vigorous.
Improved surge block design has been the subject of
some recent investigation (Schalla and Landick,
1985).
In more productive aquifers, "overpumping" was and
is a popular method for well development. With this
method, a pump is alternately turned on (usually at a
slightly higher rate than the well can sustain) and off
to simulate a surging action in the well. A problem
with this method is that the outward movement of
water normally created during surging efforts is not as
pronounced with overpumping. This may tend to
bridge the fine and coarse materials, limiting the
movement of the fine materials into the well and
thereby limiting the effectiveness of the method.
Pumping with air has also been used effectively
(Figure 5-6). Better development has been
accomplished by attaching differently shaped devices
to the end of an airline to force the air out into the
formation. An example of such a device is shown in
Figure 5-7. Such a device causes a much more
vigorous action on the movement of material in
proximity to the well screen while also pushing water
to the ground surface.
Air development techniques such as this may expose
field crews to hazardous constituents when badly
contaminated ground water is present. The technique
may also cause chemical reactions with species
present in the ground water, especially volatile
organic compounds. Care must also be taken to filter
the injected air to prevent contamination of the well
environment with oil and other lubricants present in
the compressor and airlines.
Development procedures for monitoring wells in
relatively unproductive geologic materials is
somewhat limited. Due to the low hydraulic
conductivity of the materials, surging of water in and
out of the well casing is extremely difficult. Also,
when the well is pumped, the entry rate of the water
is inadequate to effectively remove fines from the well
bore and the gravel pack material outside the well
screen.
In this type of geologic setting, where an open
borehole can be sustained, clean water can be
circulated down the well casing, out through the
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Figure 5-6 Well developments with compressed air.
Compressed Air
0 . O ' ;f I:
• o
o .
• o '
Air Development
Figure 5-7 The effects of high-velocity Jetting used for
well development through openings in a
continuous-slot well screen.
Continuous Slot Well Screen
-Jet Nozzle
Figure 5-8 Well development by back-flushing with
water.
screen, and back up the borehole (Figure 5-8).
Relatively high water velocities can be maintained and
the mud cake from the borehole wall can be broken
down effectively and removed. Because of the low
hydraulic conductivity of the geologic materials
outside the well, only a small amount of water will
penetrate the formation being monitored. This
procedure can be done before and after placement of
a gravel pack but must be conducted before a well
seal has been placed. After the gravel pack has been
placed, water should not be circulated too quickly or
the gravel pack will be lifted out of the borehole as
well. Immediately following development, the well
should be sealed, backfilled, and pumped for a short
period to stabilize the formation around the outside of
the screen and to ensure that the well will produce
fairly clear water.
5.2.7 Security
For most monitoring well installations, some
precautions must be exercised to protect the surface
portions of the well from damage. In many instances,
inadvertent vehicular accidents do occur. Monitoring
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well installations seem particularly vulnerable to
transgressions from grass mowers. Vandalism is often
a major concern, from spontaneous "hunters" looking
for a likely target to premeditated destruction of
property associated with an unpopular operation.
There are several simple solutions that can be
employed to help minimize the damage due to
accidental collisions. However, outwitting the
determined vandal may be an impossible undertaking
and certainly an expensive one.
The most basic problem to maintaining the physical
condition of any monitoring well is being able to
anticipate the hazards that might befall that particular
installation. Some instances may call for making the
well obvious to see whereas other instances may call
for keeping the well inconspicuous.
Where the most likely problem is one of vehicular
contact, be it mowers, construction traffic, or other
types of two-, three-, or four-wheeled traffic, the
first thing that can be done is to make the top of the
well plain to see. Make sure it extends far enough
above ground to be visible above grass, weeds, or
small shrubs. If that is not practical, use a "flag" that
extends above the well casing. This is also helpful for
periods when leaves or snow have buried low-lying
objects.
Paint the well casing a bright color (orange and yellow
are the most visible). This not only makes the well
more visible but also protects metal casing material
from rusting. Care should be taken to make sure paint
is not allowed inside the well casing or in threaded
fittings that may contact sampling equipment.
Make sure the owners/operators of the site being
monitored know where each installation is. Issue
maps clearly and exactly indicating where the wells
are located. Make certain their employees know the
importance of those installations, the cost associated
with them, and the difficulty involved in replacement.
The portions of the well that protrude from the ground
can also be reinforced, particularly when the well is
constructed of PVC or Teflon. The well could be
constructed such that only the portion of the well
above the water table is metal. In this manner, the
integrity of the sample is maintained as ground water
contacts only inert material and the physical condition
of the well is maintained as the upper metal portion is
better able to withstand impact.
There are two arguments to consider when
constructing a well in this manner. The arguments are
focused on the weak point in the well construction: at
or near the juncture of the metal and nonmetal
casings. One argument suggests that a longer section
of metal casing is superior because the additional
length of metal casing in the ground gives additional
strength. This way a break is less likely to occur
(although the casing is likely to be bent). The other
argument suggests that should a break in the casing
occur, a shorter length of metal casing is superior
because a break nearer to ground surface is easier to
repair. Each argument has its merits; only experience
with site conditions is likely to produce the best
solution.
The use of "well protectors" is another popular
solution that involves the use of a larger diameter
steel casing placed around the monitoring well at the
ground surface and extending several feet below
ground (Figure 5-9). The protectors are usually
seated in the cement surface seal to a depth below
the frost line.
Figure 5-9 Typical well protector installation.
Monitoring Well
_Well Protector with
Lockable Cap
Well protectors are commonly equipped with a locking
cap which insures against tampering with the inside of
the well. Dropping objects down the well can create
two potential problems: 1) impair the sampleability of
the well by clogging the well screen or impeding the
ability of the sampling device to reach water, and 2)
altering the quality of the ground water, particularly
where small quantities (perhaps drops) of an organic
liquid may be sufficient to completely contaminate the
well.
Problems associated with vandalism run from simple
curiosity to outright wanton destruction. Obviously,
sites within secured, fenced areas are less likely to
be vandalized. However, there is probably no way to
deter the determined vandal, short of posting a 24-
hour guard. In such situations, well protectors are a
must. The wells should be kept as inconspicuous as
possible. However, the benefits of "hiding" monitoring
wells must be weighed against the costs of delays in
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finding them for sampling and the potential costs for
repairs or maintenance on untried security designs.
In some situations, it might be a good policy to notify
the public of the need for the monitoring wells. If it is
properly asserted that each well serves an
environmental monitoring purpose and that the wells
have been constructed to insure public well-being, it
may create a civic conscience that would help
minimize vandalism.
As with all the previously mentioned monitoring well
components, no singular solution will best meet every
different monitoring situation. Knowledge of the social,
political, and economic conditions of the geographic
area and circumstances surrounding the need for
ground-water monitoring will dictate, to a large
degree, the type of well protection needed.
5.3 Monitoring Well Drilling Methods
As might be expected, different drilling techniques
can influence the quality of the ground-water sample
produced from a particular formation in different ways.
This applies to the drilling method employed (e.g.,
augered, driven, or rotary) as well as the driller. There
is no substitute for a conscientious driller willing to
take the extra time and care necessary to complete a
good monitoring well installation.
Among the criteria used to select an appropriate
drilling method are the following factors, listed in
order of importance:
1) Hydrologic information
a. type of formation
b. depth of drilling
c. depth of desired screen setting below top of
zone of saturation
2) Types of pollutants expected
3) Location of drilling site, i.e., accessibility
4) Design of monitoring well desired
5) Availability of drilling equipment
Table 5-1 summarizes several different drilling
methods, their advantages and their disadvantages
when used for monitoring well construction. Several
excellent publications are referenced for detailed
discussions (Campbell and Lehr, 1973; Fenn et a/.,
1977; Johnson, Inc., 1972; and Scalf et a/., 1981).
The table also gives a concept of the advantages and
disadvantages which need to be considered when
choosing a drilling technique for different site and
monitoring situations (see also, Lewis, 1982; Luhdorff
and Scalmanini, 1982; Minning, 1982; and Voytek,
1983).
Hollow- and solid-stem augering is one of the
most desirable drilling methods for constructing
monitoring wells. No drilling fluids are used and
disturbance to the geologic materials penetrated is
minimal. Auger rigs are not typically used when
consolidated rock must be penetrated and depths are
usually limited to no more than 150 feet.
In formations where the borehole will not stand open,
the monitoring well can be constructed inside the
hollow-stem augers prior to removal from the hole.
Generally, this limits the diameter of the well that can
be built to 4 inches. The hollow-stem has an added
advantage in offering the ability to collect continuous
in situ geologic samples without removal of the auger
sections.
The use of the solid-stem is most useful in fine-
grained, unconsolidated materials that will not
collapse when unsupported. The method is similar to
the hollow-stem except that the augers must be
removed from the hole to allow the insertion of the
well casing and screen. Geologic cores cannot be
collected when using a solid-stem. Therefore,
geologic sampling must rely on cuttings which come
to the surface, an undesirable method as the depth
from which the cuttings come is not precisely known.
Cable-tool drilling is one of the oldest methods used
in the water well industry. Even though the rate of
penetration is rather slow, this method offers many
advantages for monitoring well construction. With the
cable-tool, excellent formation samples can be
collected and the presence of thin permeable zones
can be detected. As drilling progresses, a casing is
normally driven and this provides an excellent
temporary casing within which the monitoring well can
be constructed.
In air-rotary drilling, air is forced down the drill stem
and back up the borehole to remove the cuttings.
This technique has been found to be particularly well
suited to drilling in fractured rock formations. If the
monitoring is intended for organic compounds, the air
must be filtered to insure that oil from the air
compressor is not introduced to the formation to be
monitored. Air-rotary should not be used in highly
contaminated environments because the water and
cuttings blown out of the hole are difficult to control
and can pose a hazard to the drill crew and
observers. Where volatile compounds are of interest,
air-rotary can volatilize those compounds and cause
water samples withdrawn from the hole to be
unrepresentative of in situ conditions. The use of
foam additives to aid cuttings removal presents the
opportunity for organic contamination of the
monitoring well.
Air-rotary with percussion hammer increases the
effectiveness of air-rotary for cavey or highly
creviced formations. Addition of the percussion
hammer gives air-rotary the ability to drive casing,
cutting the loss of air circulation in fractured rock and
maintaining an open hole in soft formations. The
capability of constructing monitoring wells inside the
driven casing prior to its being pulled adds to the
appeal of air-percussion. However, the problems
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Table 5-1 Advantages and Disadvantages of Selected Drilling Methods for Monitoring Well Construction.
Method
Drilling Principle
Advantages
Disadvantages
Drive Point
Auger, Hollow-
and Solid-stem
Jetting
Cable-tool
(Percussion)
1.25 to 2 inch ID casing with
pointed screen mechanically
depth.
Successive 5-foot flights of spiral-
shaped drill stem are rotated into
the ground to create a hole.
Cuttings are brought to the
surface by the turning action of
the auger.
Washing action of water forced
out of the bottom of the drill rod
clears hole to allow penetration.
Cuttings brought to surface by
water flowing up the outside of
the drill rod.
Hole created by dropping a heavy
"string" of drill tools into well
bore, crusing materials at bottom.
Cuttings are removed occasionally
by bailer. Generally, casing is
driven just ahead of the bottom of
the hole; a hole greater than 6
inches in diameter is usually
made.
Inexpensive.
Easy to install, by hand if
necessary.
Water samples can be collected
as driving proceeds.
Depending on overburden, a good
seal between casing and
formation can be achieved.
Inexpensive.
Fairly simple operation. Small rigs
can get to difficult-to-reach areas.
Quick set-up time.
Can quickly construct shallow
wells in firm, noncavey materials.
No drilling fluid required.
Use of hollow-stem augers greatly
facilitates collection of split-spoon
samples.
Small-diameter wells can be built
inside hollow-stem flights when
geologic materials are cavey.
Inexpensive. Driller often not
needed for shallow holes.
In firm, noncavey deposits where
hole will stand open, well
construction fairly simple.
Can be used in rock formations as
well as unconsolidated
formations.
Fairly accurate logs can be pre-
pared from cuttings if collected
often enough.
Driving a casing ahead of hole
minimizes cross-contamination by
vertical leakage of formation
waters.
Core samples can be obtained
easily.
Difficult to sample from smaller diameter
drive points if water level is below suction
lift. Bailing possible.
No formation samples can be collected.
Limited to fairly soft materials. Hard to
penetrate compact, gravelly materials.
Hard to develop. Eicreen may become
clogged if thick clays are penetrated.
PVC and Teflon® casing and screen are
not strong enough to be driven. Must use
metal construction materials which may
influence some water quality deter-
minations.
Depth of penetration limited, especially in
cavey materials. Maximum depths 150
feet.
Cannot be used in rock or well-cemented
formations. Difficult to drill in cobbles/
boulders.
Log of well is difficult to interpret without
collection of split spoons due to the lag
time for cuttings to reach ground surface.
Vertical leakage of water through borehole
during drilling is likely to occur.
Solid-stem limited to fine grained, uncon-
solidated materials that will not collapse
when unsupported.
With hollow-stem flights, heaving
materials can present a problem. May
need to add water down auger to control
heaving or wash materials from auger
before completing well.
Somewhat slow, especially with increasing
depth.
Extremely difficult to use in very coarse
materials, i.e., cobbles/boulders.
A water supply is needed that is under
enough pressure to penetrate the geologic
materials present.
Difficult to interpret sequence of geologic
materials from cuttings.
Maximum depth 150 feet, depending on
geology and water pressure capabilities.
Requires an experienced driller.
Heavy steel drive pipe used to keep hole
open and drilling "tools" can limit
accessibility.
Cannot run some geophysical logs due to
presence of drive pipe.
Relatively slow drilling method.
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Table 5-1 (continued)
Method
Drilling Principle
Advantages
Disadvantages
Hydraulic Rotary
Reverse Rotary
Air Rotary
Air-Percussion
Rotary or
Downhole-
Hammer
Rotating bit breaks formation;
cuttings are brought to the
surface by a circulating fluid
(mud). Mud is forced down the
interior of the drill stem, out the
bit, and up the annulus between
the drill stem and hole wall.
Cuttings are removed by settling
in a "mud pit" at the ground
surface and the mud is circulated
back down the drill stem.
Similar to Hydraulic Rotary
method except the drilling fluid is
circulated down the borehole out-
side the drill stem and is pumped
up the inside, just the reverse of
the normal rotary method. Water
is used as the drilling fluid, rather
than a mud, and the hole is kept
open by the hydrostatic pressure
of the water standing in the bore-
hole.
Very similar to Hydraulic Rotary,
the main difference being that air
is used as the primary drilling fluid
as opposed to mud or water.
Air Rotary with a reciprocating
hammer connected to the bit to
fracture rock.
Drilling is fairly quick in all types
of geologic materials.
Borehole will stay open from
formation of a mud wall on sides
of borehole by the circulating
drilling mud. Eases geophysical
logging and well construction.
Geologic cores can be collected.
Virtually unlimited depths
possible.
Creates a very "clean" hole, not
dirtied with drilling mud.
Can be used in all geologic
formations.
Very deep penetrations possible.
Split-spoon sampling possible.
Can be used in all geologic forma-
tions; most successful in highly
fractured environments.
Useful at any depth.
Fairly quick.
Drilling mud or water not
required.
Very fast penetrations.
Useful in all geologic formations.
Only small amounts of water
needed for dust and bit tempera-
ture control.
Cross-contamination potential can
be reduced by driving casing.
Expensive, requires experienced driller and
fair amount of peripheral equipment.
Completed well may be difficult to
develop, especially small-diameter wells,
because of mud wall on borehole.
Geologic logging by visual inspection of
cuttings is fair due to presence of drilling
mud. Thin beds of sand, gravel, or clay
maybe missed.
Presence of drilling mud can contaminate
water samples, especially the organic, bio-
degradable muds.
Circulation of drilling fluid through a
contaminated zone can create a hazard at
the ground surface with the mud pit and
cross-contaminate clean zones during
circulation.
A large water supply is needed to maintain
hydrostatic pressure in deep holes and
when highly conductive formations are
encountered.
Expensive—experienced driller and much
peripheral equipment required.
Hole diameters are usually large,
commonly 18 inches or greater.
Cross-contamination from circulating
water likely.
Geologic samples brough to surface are
generally poor, circulating water will
"wash" finer materials from sample.
Relatively expensive.
Cross-contamination from vertical
communication possible.
Air will be mixed with water in the hole
and that which is blown from the hole,
potentially creating unwanted reactions
with contaminants; may affect
"representative" samples.
Cuttings and water blown from the hole
can pose a hazard to crew and surrounding
environment if toxic compounds
encountered.
Organic foam additives to aid cuttings
removal may contaminate samples.
Relatively expensive.
As with most hydraulic rotary methods,
the rig is fairly heavy, limiting accessibility.
Vertical mixing of water and air creates
cross-contamination potential.
Hazard posed to surface environment if
toxic compounds encountered.
Organic foam additives for cuttings
removal may contaminate samples.
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with contamination and crew safety must still be
considered.
Reverse-rotary drilling has limited application for
monitoring well construction. Reverse-rotary
requires that large quantities of water be circulated
down the borehole and up the drill stem to remove
cuttings. If permeable formations are encountered,
significant quantities of water can move into the
formation to be monitored, altering the quality of the
water to be sampled.
Hydraulic rotary, or "mud" rotary, is probably the
most popular method used in the water well industry.
However, hydraulic rotary presents some
disadvantages for monitoring well construction. With
hydraulic rotary, a drilling mud (usually bentonite) is
circulated down the drill stem and up the borehole to
remove cuttings. The mud creates a wall on the side
of the borehole which must be removed from the
screened area by development procedures. With
small diameter wells, complete removal of the drilling
mud is not always achieved. The ion exchange
potential of most drilling muds is high and may
effectively reduce the concentration of trace metals in
water entering the well. In addition, the use of
biodegradable, organic drilling muds can introduce
organic components to water sampled from the well.
Most ground-water monitoring wells will be
completed in glaciated or unconsolidated materials
and will be relatively shallow, perhaps less than 50 to
75 feet. In these applications, hollow-stem augering
usually will be the method of choice. Solid-stem
auger, cable-tool, and air-percussion also offer
advantages depending on the geology and
contaminant of interest.
5.3.7 Geologic Samples
Permit applications for disposal of waste materials
often require that geologic samples be collected at
the disposal site. Investigations of ground-water
movement and contaminant transport should also
include the collection of geologic samples for physical
inspection and testing. Opportunity for stratigraphic
sample collection is best afforded during monitoring
well drilling.
Samples can be collected continuously, at each
change in stratigraphic unit, or, in homogeneous
materials, at regular intervals. These samples may
later be classified, tested, and analyzed for physical
properties such as particle size distribution, textural
classification, and hydraulic conductivity, and for
chemical analyses such as ion-exchange capacity,
chemical composition, and specific parameter
teachability.
Probably the most common method of material
sampling is with a "split-spoon" sampler. This
device is a 12- or 18-inch long hollow cylinder (2-
inch diameter) which is split in half lengthwise. The
halves are held together at each end with threaded
couplings; the top end attaches to the drill rod, and
the bottom end is a drive shoe (F:igure 5-lOa). The
sampler is lowered to the bottom of the hole and
driven ahead of the hole with a weighted "hammer"
striking an anvil at the upper end of the drill rod to
which the sampler is attached. Sample is forced up
the inside of the tube and is held with a basket trap or
flap valve that allow the sample to enter the sampler
but not exit (though retention of noncohesive, sandy
formations is often difficult). After the sampler is
withdrawn from the hole, the sample is removed by
unscrewing the ends and separating the sample
collection tube.
Another common sampler is the thin wall tube or
"Shelby" tube. These tubes are usually 2 to 5-1/2
inches in diameter and about 24 inches long. The
cutting edge of the tube is sharpened and the upper
end is attached to a coupling head by means of cap
screws or a retaining pin (Figure 5-1 Ob). A Shelby
tube has a minimum ratio of wall area to sample area
and creates the least disturbance to the sample of
any drive-type sampler in current use (for hydraulic
conductivity tests of low conductivity, <10"6 cm/sec
materials, minimal disturbance is critical). After
retraction, the tube is disconnected from the head
and the sample is removed from the tube with a jack
or press. If sample preservation is a major concern,
the tube can be sealed and shipped to the laboratory.
Apart from permit requirements, material samples are
very helpful for deciding at what depth to complete a
monitoring well. Unexpected changes encountered
during drilling can alter preconceived ideas
concerning the local ground-water flow regime. In
many instances, the driller will be able to detect a
change in formation by a change in penetration rate,
sound, or "feel" of the drilling rig. However, due to
the lag time for cuttings to come to the surface and
the amount of mixing the cuttings may undergo as
they come up the borehole, the only way to truly
know what the subsurface materials look like is to
stop drilling and collect a sample.
5.3.2 Case History
Several different types of monitoring wells were
constructed during the investigation of a volatile
organic contaminant plume in northern Illinois
(Wehrmann, 1984). A brief summary of the types of
wells employed and the reasons for their use help
illustrate how an actual ground-water quality
monitoring problem was approached.
During the final weeks of a one-year study of
ground-water nitrate quality in north central Illinois,
the presence of a number of organic compounds was
detected in the drinking water of all five homes
sampled within a large rural residential subdivision.
The principal compound found was trichloroethylene,
TCE, at concentrations between 50 and 1,000
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Figure 5-10 Crass-sectional views of (a) split spoon and
(b) Shelby tube samplers (from Mobile
Drilling Co., 1972).
Split Spoon
Shelby Tube
micrograms per liter (ug/l). All the homes in the
subdivision utilized private wells tapping a surficial
sand and gravel deposit at a depth of 65 to 75 feet. A
geologic cross-section of the study area is shown in
Figure 5-11.
Two immediate concerns needed to be addressed.
First, how many other drinking wells were affected
and, second, what was the contaminant source? Early
thoughts connected the TCE contaminant to the
contamination potential of the large number of septic
systems in the subdivision. Previous work
(Wehrmann, 1983) had established the ground-water
flow direction beneath the affected subdivision was
from north-northeast to south-southwest. Because
the area upgradient of the subdivision was primarily
farmland, several monitoring wells placed upgradient
of the subdivision would help confirm or deny the
possibility that the septic systems were the source of
the VOC contamination.
Five "temporary" monitoring wells were constructed
in the field upgradient of the affected subdivision.
Original plans called for driving a 2-inch diameter
sandpoint to depths from 40 to 70 feet. Samples
would be collected at 10-foot intervals as the point
was driven. Once 70 feet was reached, the sandpoint
would be pulled, the hole properly abandoned, and
the point driven at a new sampling location. The first
hole was to be placed north (upgradient) of a
domestic well found to be highly contaminated. Holes
were to be placed successively in an upgradient
direction proceeding across the field. In this manner,
ground-water samples could be quickly collected at
many depths and locations, the well materials
recovered, and the field left relatively undisturbed.
Once drilling commenced, however, it became clear
that it was not possible to drive 2-inch sandpoints
into the coarse sand and gravel just below ground
surface in this area. An air-percussion rig was
brought on-site and a new approach was used. A
4-inch diameter screen (2 feet long) with a drive
shoe was welded to a 4-inch diameter steel casing.
This assembly was driven by air hammer to the
desired sampling depth. The bottom of the drive
shoe, being open, forced the geologic materials
penetrated into the casing and screen. These
materials were evacuated from the casing and screen
by air rotary once the desired depth was reached. To
avoid cross-contamination from using the same
materials at several locations, all well materials were
steam cleaned prior to use and between holes.
Locations of the temporary well sites and the
analytical results for TCE from samples bailed at
depths of 40 and 50 feet are shown in Figure 5-12.
Results of the temporary sampling revealed the
contaminant source was indeed outside of the
subdivision. Due to the construction and sampling
methods employed for these wells, emphasis was not
placed on the quantitative aspects of the sampling
results. However, important qualitative conclusions
were made. The temporary wells confirmed the
presence of VOCs directly upgradient of the
subdivisions and provided information for the location
and depth of nine permanent monitoring wells.
Due to the problems associated with organic
compound teachability and adsorption from PVC
casing and screen, flush-threaded stainless steel
casing and screen, 2-inches in diameter, were used
for the permanent sampling wells. The screens were
2 feet long with 0.01-inch wire-wound slot
openings. All materials associated with the monitoring
well construction, including the drill rig, were steam
cleaned prior to the commencement of drilling to
avoid organic contamination from cutting oils and
grease. Prior to use, the casing and screen materials
were kept off-site in a covered, protected area.
To insure that the sandy materials would not collapse
the hole after drilling, casing lengths and the screen
were screwed together above ground and placed
down the inside of the augers before the auger flights
were pulled out of the hole. The sand and gravel
below the water table collapsed around the screen
and casing as the augers were removed. To help
prevent vertical movement of water down along the
casing, a wet bentonite/cement mixture was placed in
99
-------
Figure 5-11 East-west cross section across Rock River Valley at Roscoe (from Berg et al., 1981).
Rock River Floodplain
Galena-Platteville Dolomite
Glenwood-St. Peter Sandstone
:<>:>:] Till
':;V •'.::] Outwash sand and gravel vT;,-:.','/.','»-[•.•:-.I' ;• '•• •
i u''iVr>l Lacustrine sands, silt and
Organic materials (or) buried soil
WW - Water Well
T - Tollway boring
500
Scale (miles)
the annulus just above the water table to a thickness
of 2 to 3 feet. Cuttings (principally clean, medium to
fine sand) were backfilled above the bentonite/cement
seal to within 4 feet of land surface. Another
bentonite/cement mixture was placed to form a seal
at ground surface, further preventing movement of
water down along the well casing. A 4-inch diameter
steel protective cover with locking cap was placed
around the protruding casing and into the surface seal
to protect against vandalism.
The nine wells were drilled at four locations with
paired wells at three sites and a nest of three wells at
one site (Figure 5-13). The locations were based on
the analytical results of the samples taken from the
temporary wells and basic knowledge of the ground-
water flow direction. Locations were numbered as
nests 1 through 4 in order of their construction. Nest
1, located immediately north of the affected
subdivision, consists of three wells completed at
approximately 60, 70 and 80 feet below ground
surface. Nest 2 consists of two wells 50 and 60 feet
deep. Nest 3 consists of two wells constructed to 40
and 55 feet and nest 4 consists of two wells
completed at 50 and 60 feet.
Subsequent to the completion of these nine wells, it
was felt an additional well constructed to 100 feet at
the location of nest 1 was needed to further define
the vertical extent of the contaminant plume. Because
the hollow-stem auger rig was no longer available,
arrangements were made to use a cable-tool rig to
drill the hole. The well was constructed over a period
of two days, somewhat slower than any of the other
methods previously used (but typical of cable-tool
speeds). With this method, a 6-inch casing was
driven several feet, a bit was used to break up the
materials inside the casing, then the materials were
removed from the casing with a dart-valve bailer.
This procedure was repeated until the desired depth
of 100 feet was reached. Once this depth was
reached, the well casing and screen were screwed
together and lowered down the hole. The 6-inch
casing was then pulled back allowing the hole to
collapse about the well (the well was constructed of
stainless steel exactly as the nine other monitoring
wells). As before, all drilling equipment and well
construction materials were steam cleaned prior to
use.
Appraisal of the results of sampling these monitoring
wells and the domestic wells in the area produced the
pictorial representations shown in Figures 5-14 and
5-15. Figure 5-14 conceptually illustrates a cross
section of the TCE plume looking in the general
direction of ground-water flow in the vicinity of
monitoring nests 2, 3, and 4. The likely extent of the
VOC contaminant plume is shown in Figure 5-15.
This map includes a limited amount of data from
privately owned monitoring wells located on industrial
property just upgradient of monitoring nests 2 and 4.
100
-------
Figure 5-12 Locations and TCE concentrations for temporary monitoring wells at Roscoe, Illinois (from Wehrmann, 1984).
EXPLANATION
ppb TCE @ 40'
TEMPORARY
WELL LOCATION
ppb TCE @ 50'
HOUSE WELL
SAMPLED
SAME WEEK,
2178 ppb TCE @ 65
HONONEGAH
COUNTRY ESTATES
VILLAGE OF
ROSCOE
MOORE HAVEN
SUBDIVISION
101
-------
Figure 5-13 Location of monitoring well nests and cross-section A-A'at Roscoe, Illinois (from Wehrmann, 1984).
SCALE OF'FSET
500 1QOO 15OO JOOO
NEST 2
50', 60'
NEST 4
50', 60'
NEST 1
60', 70', 80
HONONEGAH
COUNTRY ESTATES
VILLAGE OF
ROSCOE
i i
MOORE HAVENl
SUBDIVISION
i \, MNHIKINNICK |J5HOOL ]
\ q----re"rm
102
-------
The dashed lines indicate the probable extent of the
contaminant plume based on the dimensions of the
plume as it passes beneath the developed area along
the Rock River.
This monitoring situation clearly indicates the role
different drilling and construction techniques can take
in a ground-water sampling strategy. In each
instance, much consideration was given to the effect
the methods used for construction and sampling
would have on the resultant chemical data. Where
quantitative results for a fairly "quick" preliminary
investigation were not necessary and, after
determining that it was too difficult to drive
sandpoints, it was felt using air-percussion rotary
was acceptable. For the placement of the permanent
monitoring wells, wells that may become crucial for
contaminant source identification and possibly be
involved in litigation, the hollow-stem auger was the
technique of choice. Finally, when the hollow-stem
auger was not available and it was decided another
hole was needed, the cable-tool rig was chosen.
Here, it was recognized that only one hole was to be
drilled so the relative slowness of the method became
less a factor. Also, the depth of completion (100 feet)
in the cavey sand and gravel made cable-tool
preferable over the hollow-stem. Note, too, that
each method chosen was capable of maintaining an
open hole without the use of drilling mud which could
have affected the results of the organic compound
analyses.
5.4 Summary
Critical considerations for the design of ground-
water quality monitoring networks include alternatives
for well design and drilling techniques. With a
knowledge of the principal chemical constituents of
interest, local hydrogeology, and an appreciation of
subsurface geochemistry, appropriate selections of
materials for well design and drilling techniques can
be made. Whenever possible, physical disturbance
and the amount of foreign material introduced into the
subsurface should be minimized.
The choices of drilling methods and well construction
materials are very important decisions to be made in
every type of ground-water monitoring program.
Details of network construction can introduce
significant bias into monitoring data which frequently
may be corrected only by repeating the process of
well siting, installation, completion, and development.
This can be quite costly in time, effort, money, and
loss of information. Undue expense is avoidable if
planning decisions are made cautiously with an eye to
the future.
The expanding scientific literature on effective
ground-water monitoring techniques should be read
and evaluated on a continuing basis. This information
will help supplement guidelines, such as this, for
applications to specific monitoring efforts.
5.5 References
Barcelona, M.J., J.P. Gibb, J.A. Helfrich, and E.E.
Garske. 1985. Practical Guide for Ground-Water
Sampling. Illinois State Water Survey. U.S.
Environmental Protection Agency, Robert S. Kerr
Environmental Research Laboratory, Ada, OK and
Environmental Monitoring and Support Laboratory,
Las Vegas, NV.
Barcelona, M.J. 1984. TOG Determinations in Ground
Water. Ground Water 22(1): 18-24.
Barcelona, M.J., J.A. Helfrich, E.E. Garske, and J.P.
Gibb. 1984. A Laboratory Evaluation of Ground Water
Sampling Mechanisms. Ground Water Monitoring
Review 4(2):32-41.
Barcelona, J.J., J.P. Gibb, and R.A. Miller. 1984. A
Guide to the Selection of Materials for Monitoring
Well Construction and Ground-Water Sampling.
Illinois State Water Survey Contract Report 327.
Illinois State Water Survey, Champaign, IL.
Berg, R.C., J.P. Kempton, and A.N. Stecyk. 1981.
Geology for Planning in Boone and Winnebago
Counties, Illinois. Illinois State Geological Survey
Circular 531, Illinois State Geological Survey, Urbana,
IL.
Brown, K.W., J. Green, and J,C. Thomas. 1983. The
Influence of Selected Organic Liquids on the
Permeability of Clay Liners. Proceedings of the Ninth
Annual Research Symposium: Land Disposal,
Incineration, and Treatment of Hazardous Wastes.
U.S. Environmental Protection Agency
SHWRD/EPCS, May 2-4, 1983, Ft. Mitchell, KY.
Burkland, P.W., and E. Raber. 1983. Method to Avoid
Ground-Water Mixing Between Two Aquifers During
Drilling and Well Completion Procedures. Ground
Water Monitoring Review 3(4):48-55. Campbell,
M.D. and J.H. Lehr. 1973. Water Well Technology.
McGraw-Hill Book Company, New York, NY.
Fenn, D., E. Cocozza, J. Isbister, 0. Braids, B. Yare,
and P. Roux. 1977. Procedures Manual for Ground
Water Monitoring at Solid Waste Disposal Facilities
(SW-611). U.S. Environmental Protection Agency,
Cincinnati, OH.
Gibb, J.P., R.M. Schuller, and R.A. Griffin. 1981.
Procedures for the Collection of Representative Water
Quality Data from Monitoring Wells. Cooperative
Groundwater Report. Illinois State Water and
Geological Surveys, Champaign, IL.
Gillham, R.W., M.J.L. Robin, J.F. Barker, and J.A.
Cherry. 1983. Groundwater Monitoring and Sample
Bias. American Petroleum Institute Publication 4367,
Environmental Affairs Department.
103
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Figure 5-14 Cross-Section A-A' through monitoring nests 2, 3, and 4, looking in the direction of ground-water flow
(from Wehrmann, 1984).
Top of Saturated Zone (Water Table)
100-250^g/L ^_
<100^g/L
Illinois State Water Survey and Illinois State
Geological Survey. 1984. Proceedings of the 1984
ISWS/ISGS Groundwater Monitoring Workshop.
February 27-28, Champaign, IL.
Illinois State Water Survey and Illinois State
Geological Survey. 1982. Proceedings of the 1982
ISWS/ISGS Groundwater Monitoring Workshop.
Illinois Section of American Water Works Association.
February 22-23, 1982, Champaign, IL.
Johnson, T.L. 1983. A Comparison of Well Nests vs.
Single-Well Completions. Ground Water Monitoring
Review 3(1):76-78.
Johnson, E.E., Inc. 1972. Ground Water and Wells.
Johnson Division, Universal Oil Products Co., St.
Paul, MN.
Keith, S.J., L.G. Wilson, H.R. Fitch, D.M. Esposito.
1982. Sources of Spatial-Temporal Variability in
Ground-Water Quality Data and Methods of Control:
Case Study of the Cortaro Monitoring Program,
Arizona. Proceedings of the Second National
Symposium on Aquifer Restoration and Ground Water
Monitoring. National Water Well Association, May
26-28, 1982, Columbus, OH.
Lewis, R.W. 1982. Custom Designing of Monitoring
Wells for Specific Pollutants and Hydrogeologic
Conditions. Proceedings of the Second National
Symposium on Aquifer Restoration and Ground Water
Monitoring. National Water Well Association, May
26-28, 1982, Columbus, OH.
Luhdorff, E.E., Jr., and J.C. Scalmanini. 1982.
Selection of Drilling Method, Well Design and
Sampling Equipment for Wells for Monitor Organic
Contamination. Proceedings of the Second National
Symposium on Aquifer Restoration and Ground Water
Monitoring, National Water Well Association, May
26-28, 1982, Columbus, OH.
Mackay, D.M., P.V. Roberts, and J.A. Cherry. 1985.
Transport of Organic Contaminants in Groundwater.
Environmental Science & Technology 19(5):384-
392.
Miller, G.D. 1982. Uptake and Release of Lead,
Chromium and Trace Level Volatile Organics Exposed
to Synthetic Well Castings. Proceedings of Second
National Symposium on Aquifer Restoration and
Ground Water Monitoring. National Water Well
Association, May 26-28, 1982, Columbus, OH.
Minning, R.C. 1982. Monitoring Well Design and
Installation. Proceedings of the Second National
Symposium on Aquifer Restoration and Ground Water
Monitoring. National Water Well Association, May
26-28, 1982, Columbus, OH.
Mobile Drilling Company. 1972. Soil Sampling
Equipment - Accessories. Catalog 650. Mobile
Drilling Company, Indianapolis, IN.
Morrison, R.D. and P.E. Brewer. 1981. Air-Lift
Samplers for Zone of Saturation Monitoring. Ground
Water Monitoring Review 1(1):52-55.
Naymik, T.G. and M.E. Sievers. 1983. Groundwater
Tracer Experiment (II) at Sand Ridge State Forest,
Illinois. Illinois State Water Survey Contract Report
334. Illinois State Water Survey, Champaign, IL.
Naymik, T.G. and J.J. Barcelona. 1981.
Characterization of a Contaminant Plume in
Groundwater, Meredosia, Illinois. Ground Water
16(3):149-157.
104
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Figure 5-15 General area of known TCE contamination (from Wehrmann, 1984).
DOMESTIC WELLS
8/23/83
Wa/L
COMMERCIAL
OLO6 FARM/TRESEMER SUBDIVISION
21 DOMESTIC WELLS
SAMPLED 10/3 -4/83
pg/L to 4
N4 COMMERCIAL
« WELLS
L _l 3/15/83
I ^U-22
P,,
AREA KNOWN
CONTAIN
TCE
(10 pg/L to >2000 flg/L)
COMMERCIAL
WELL
6/21/83
4Llg/L
105
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O'Hearn, M. 1982. Groundwater Monitoring at the
Havana Power Station's Ash Disposal Ponds and
Treatment Lagoon. Confidential Contract Report.
Illinois State Water Survey, Champaign, IL.
Perry, C.A., and R.J. Hart. 1985. Installation of
Observation Wells on Hazardous Waste Sites in
Kansas Using a Hollow-Stem Auger. Ground
Monitoring Review 5(4):70-73.
Pettyjohn, W.A., and A.W. Hounslow. 1982. Organic
Compounds and Ground-Water Pollution.
Proceedings of the Second National Symposium on
Aquifer Restoration and Ground Water Monitoring.
National Water Well Association, May 26-28, 1982,
Columbus, OH.
Pettyjohn, W.A., W.J. Dunlap, R. Cosby, and J.W.
Keeley. 1981. Sampling Ground Water for Organic
Contaminants. Ground Water 19(2):180-189.
Pfannkuch, H.O. 1981. Problems of Monitoring
Network Design to Detect Unanticipated
Contamination. Proceedings of the First National
Ground Water quality Monitoring Symposium and
Exposition. National Water Well Association, May
29-30, 1981, Columbus, OH.
Pickens, J.F., J.A. Cherry, R.M. Coupland, G.E.
Grisak, W.F. Merritt, and B.A. Risto. 1981. A Multi-
Level Device for Ground-Water Sampling. Ground
Water Monitoring Review 1(1):48-51.
Rinaldo-Lee, M.B. 1983. Small - vs. Large-
Diameter Monitoring Wells. Ground Water Monitoring
Review 3(1):72-75.
Scalf, M.R., J.F. McNabb, W.J. Dunlap, R.L. Cosby,
and J. Fryberger. 1981. Manual of Ground-Water
Sampling Procedures. NWWA/EPA Series, National
Water Well Association, Worthington, OH.
Schalla, R., and R.W. Landick. 1985. A New Valved
and Air-Vented Surge Plunger for Developing
Small-Diameter Monitor Wells. Proceedings of the
Third National Symposium and Exposition on
Ground-Water Instrumentation. National Water Well
Association, October 2-4, 1985, San Diego, CA.
Sosebee, J.B., Jr. ef a/. 1982. Contamination of
Groundwater Samples with PVC Adhesives and PVC
Primer from Monitor Wells. Environmental Science
and Engineering, Inc., Gainesville, FL.
Torstensson, B.A. 1984. A New System for Ground
Water Monitoring. Ground Water Monitoring Review
4(4):131-138.
Voytek, J.E., Jr. 1983. Considerations in the Design
and Installation of Monitoring Wells. Ground Water
Monitoring Review 3(1):70-71.
Walker, W.H. 1974. Tube Wells, Open Wells, and
Optimum Ground-Water Resource Development.
Ground Water 12(1): 10-15.
Wehrmann, H.A. 1984. An Investigation of a Volatile
Organic Chemical Plume in Northern Winnebago
County, Illinois. Illinois State Water Survey Contract
Report 346, Illinois State Water Survey, Champaign,
IL.
Wehrmann, H.A. 1983. Monitoring Well Design and
Construction. Ground Water Age~17(8):35-38.
Wehrmann, H.A. 1983. Potential Nitrate
Contamination of Groundwater in the Roscoe Area,
Winnebago County, Illinois. Illinois State Water Survey
Contract Report 325, Illinois State Water Survey,
Champaign, IL.
Wehrmann, H.A. 1982. Groundwater Monitoring for
Fly Ash Leachate, Baldwin Power Station, Illinois
Power Company. Confidential Contract Report. Illinois
State Water Survey, Champaign, IL.
106
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CHAPTER 6
GROUND- WA TER SAMPLING
6.1 Introduction
6.1.1 Background
Ground-water sampling is conducted to provide
information on the condition of our subsurface water
resources. Whether the goal of the monitoring effort
is detection or assessment of contamination, the
information gathered during sampling efforts must be
of known quality and be well documented. The most
efficient way to accomplish these goals for water
quality information is the development of a sampling
protocol which is tailored to the information needs of
the program and the hydrogeology of the site or
region under investigation. This sampling protocol
incorporates detailed descriptions of sampling
procedures and other techniques which of themselves
are not sufficient to document data quality or
reliability. Sampling protocols are central parts of
networks or investigatory strategies.
The need for reliable ground-water sampling
procedures has been recognized for years by a
variety of professional, regulatory, public and private
groups. The technical basis for the use of selected
sampling procedures for environmental chemistry
studies has been developed for surface-water
applications over the last four decades. However,
ground-water quality monitoring programs have
unique needs and goals which are fundamentally
different from previous investigative activities. The
reliable detection and assessment of subsurface
contamination require minimal disturbance of
geochemical and hydrogeologic conditions during
sampling.
At this time proven well construction, sampling and
analytical protocols for ground-water sampling have
been developed for many of the more problematic
chemical constituents of interest. However, the
acceptance of these procedures and protocols must
await more careful documentation and firm regulatory
guidelines for monitoring program execution. The time
and expense of characterizing actual subsurface
conditions place severe restraints on the methods
which can be employed. Since the technical basis for
documented, reliable drilling, sample collection and
handling procedures is in the early stages of
development, conscientious efforts to document
method performance under real conditions should be
a part of any ground-water investigation (Barcelona
etal., 1985; Scalf efa/., 1981).
6.7.2 Information Sources
Much of the literature on routine ground-water
monitoring methodology has been published in the
last 10 years. The bulk of this work has emphasized
ambient resource or contaminant resource monitoring
(detection and assessment) rather than case
preparation or enforcement efforts. General
references which are useful to the design and
execution of sampling efforts are those of the U.S.
Geological Survey (1977; Wood, 1976), the U.S.
Environmental Protection Agency (Brass et al., 1977;
Dunlap et at., 1977; Fenn et al., 1977; Sisk, 1981)
and others (National Council of the Paper Industry,
1982; Tinlin, 1976). In large part, these past works
treat sampling in the context of overall monitoring
programs, providing descriptions of available sampling
mechanisms, sample collection and handling
procedures. The impact of specific methodologies on
the usefulness or reliability of the resulting data have
received little discussion (Gibb ef al., 1981; Todd et
al., 1976).
High quality chemical data collection is essential in
ground-water monitoring programs. The technical
difficulties involved in "representative" sampling have
been recognized only recently (Gibb ef al., 1981;
Grisak ef al., 1978). It is clear that the long-term
collection of high quality ground-water chemistry
data is more involved than merely selecting a
sampling mechanism and agreeing on sample
handling procedures. Efforts to detect and assess
contamination can be unrewarding without accurate
(e.g., unbiased) and precise (e.g., comparable and
complete) concentration data on ground-water
chemical constituents. Also, the expense of data
collection and management argue for documentation
of data quality.
Gillham ef al. (1983) have published a very useful
reference on the principal sources of bias (i.e.,
inaccuracy) and imprecision (i.e., nonreproducibility)
in ground-water monitoring results. Their treatment
is extensive and stresses the minimization of random
error which can enter into well construction, sample
107
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collection and sample handling operations.They
further stress the importance of collecting precise
data over time to maximize the effectiveness of trend
analysis, particularly for regulatory purposes.
Accuracy is also very important, since the ultimate
reliability of statistical comparisons of results from
different wells (e.g., upgradient versus downgradient
samples) may depend on differences between mean
values for selected constituents from relatively small
replicate sample sets. Therefore, systematic error
must be controlled by selecting proven methods for
establishing sampling points and sample collection to
insure known levels of accuracy.
6.7.3 The Subsurface Environment
The subsurface environment of ground water may be
categorized broadly into two zones, the unsaturated
(i.e., vadose) and saturated zones. The use of the
term vadose zone is more accurate as isolated
water-saturated regions may exist in the unsaturated
zone above the table or most shallow confined
aquifer.
Scientists and engineers have discovered recently
that the subsurface is neither devoid of oxygen
(Winograd and Robertson, 1982) nor sterile (Wilson
and McNabb, 1983; Wilson et a/., 1983). These facts
may have significant influence on the mobility and
persistence of chemical species, as well as on the
transformations of the original components of
contaminant mixtures (Schwarzenbach et a/., 1985)
which have been released to the subsurface.
The subsurface environment is also quite different
from surface water systems in that vertical gradients
in pressure and dissolved gas content have been
observed within the usual depth ranges of monitoring
interest (i.e., 1 to 150 m [3 to 500 ft]). These
gradients can be linked to well-defined hydrologic or
geochemical processes in some cases. However,
reports of apparently anomalous geochemical
processes have increased in recent years, particularly
at contaminated sites (Barcelona and Garske, 1983;
Heaton and Vogel, 1981; Schwarzenbach et a/., 1985;
Winograd and Robertson, 1982; Wood and Petraitis,
1984).
The subsurface environment is not as readily
accessible as surface water systems, and some
disturbance is necessary to collect samples of earth
materials or ground water. Therefore,
"representative" (i.e., artifact or error free) sampling
is really a function of the degree of detail needed to
characterize subsurface hydrologic and geochemical
conditions and the care taken to minimize disturbance
of these conditions in the process (Claasen, 1982).
Each well or boring represents a potential conduit for
short-circuited contaminant migration or ground-
water flow which must be considered a potential
liability to investigative activities.
It is clear that the subsurface environment of ground
water is dynamic over extended time frames and the
processes of recharge and ground-water flow are
very important to a thorough understanding of the
system. Detailed descriptions of contaminant
distribution, transport and transformation necessarily
rely on the understanding of basic flow and fluid
transport processes. It is important to keep in mind
that short-term investigations may only provide a
snapshot of contaminant levels or distributions. Since
water quality monitoring data is normally collected on
discrete dates, it is very important that reliable
collection methods are used which assure high data
quality over the course of the investigation. The
reliability of the methods should be investigated
thoroughly during the preliminary phase of monitoring
network implementation.
Though the scope of this discussion is on sampling
ground water for chemical analysis, it should be
emphasized that the same data quality requirements
apply to water level measurements and to hydraulic
conductivity testing. These hydrologic determinations
are the basis for the interpretation of chemical
constituent data and may well limit the validity of fluid
or solute transport model applications. Hydrologic
measurements must be included in the development
of the quality assurance/quality control program for
ground-water quality monitoring networks.
6.7.4 The Sampling Problem and Parameter
Selection
Cost-effective water quality sampling is difficult in
ground-water systems, because proven field
procedures have not been extensively documented.
Regulations which call for "representative sampling"
alone are not sufficient to insure high quality data
collection. The most appropriate monitoring and
sampling procedures for a ground-water quality
network will depend on the specific purpose of the
program. Resource evaluation, contaminant detection,
remedial action assessments and litigation studies are
purposes for which effective networks can be
designed once the information needs have been
identified. Due to the time, manpower and cost of
most water-quality monitoring programs, the optimal
network design should be phased so as to make the
most of the available information as it is collected.
This approach allows for the gradual refinement of
program goals as the network is implemented.
There are two fundamental considerations which are
common to most ground-water quality monitoring
programs. These are establishing individual sampling
points (i.e., in space and time) and the elements of
the water sampling protocol which will be sufficient to
meet the information needs of the overall program.
The placement and number of sampling points can be
phased to gradually increase the scale of the
monitoring program. Similarly, the chemical
constituents of initial interest should provide
108
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background ground-water quality data from which a
list of likely contaminants may be prepared as the
program progresses. Candidate chemical and
hydrologic parameters for both detective and
assessment monitoring activities are shown in Table
6-1. Special care should be taken to account for
possible subsurface transformation of the principal
pollutant species. Ground-water transport of
contaminants can produce chemical distributions
which vary substantially over time and space.
Transformation of organic compounds in particular
can change the identity of the original contaminant
mixture substantially (Mackay et al., 1985;
Schwarzenbach ef a/., 1985).
Table 6-1 Suggested Measurements for Ground-Water
Monitoring Programs
Detective Monitoring
Chemical Parameters"
pH, Q-', TOC, TOX, Alkalinity, TDS, Eh, C|-, NCV, SO/, P04S,
Si02, Na*, K*, Ca", Mg", NH/, Fe, Mn
Hydrologic Parameters
Water Level, Hydraulic Conductivity
Assessment Monitoring
Chemical Parameters"
pH, Q-, TOC, TOX, Alkalinity, TDS, Eh, CI", N03-, S04=, P04S,
Si02, B, Na*, K*, Ca", Mg", NH4*, Fe, Mn, Zn, Cd, Cu, Pb, Cr,
Ni, Ag, Hg, As, Sb, Se, Be
Hydrologic Parameters
Water Level, Hydraulic Conductivity
°Q~' = specific conductance, a measure of the charged species in
solution.
Source: Barcelona et al., 1981.
Contaminant detection is generally the most important
aspect of a water quality program which must be
assured in network design. False negative
contaminant detections due to the loss of chemical
constituents or the introduction of interfering
substances which mask the presence of the
contaminants in water samples can be very serious.
Such errors may delay needed remedial action and
expose either the public or the environment to
unreasonably high risk. False positive observations of
contaminants may call for costly remedial actions or
more intensive study which are not warranted by the
actual situation. Reliable sample collection and data
interpretation procedures are therefore central to an
optimized network design.
In this respect, monitoring in the vadose zone is
attractive because it should provide an element of
"early" detection capability. The methodologies
available for this type of monitoring have been under
development for some time. However, there are
distinct limitations to many of the available monitoring
devices (Everett and McMillan, 1985; Everett ef al.,
1982; Wilson, 1981; Wilson, 1982; Wilson, 1983) and
it is frequently difficult to relate observed vadose-
zone concentrations quantitatively to actual
contaminant distributions in ground water (Everett ef
a/., 1984; Lindau and Spalding, 1984). Soil gas
sampling techniques and underground storage tank
monitors have been commercially developed,
however, which can be extremely useful for source
scouting. Given the complexity of vadose zone
monitoring procedures and the need for additional
investigation (Bobbins and Gemmell, 1985), it may be
difficult to implement these techniques in routine
ground-water monitoring networks.
This chapter addresses water quality sampling in the
saturated zone, reflecting the advanced state of
monitoring technology appropriate for this
compartment of the subsurface. There are a number
of useful reference materials for the development of
effective ground-water sampling protocols which
include information on the types of drilling methods,
well construction materials, sampling mechanisms
and sample handling methods currently available
(Barcelona ef al., 1985; Barcelona ef al., 1983;
Gillham ef al., 1983; Scalf ef al., 1981; Todd ef a/.,
1976). In order to collect sensitive, high-quality
contaminant concentration data, it is important to
identify the type and magnitude of errors which may
arise in ground-water sampling. A generalized
diagram of the steps involved in sampling and
principal sources of error is shown in Figure 6-1. It
should be recognized that strict error control at each
step is necessary for the collection of high quality
data which is representative of the in situ condition.
Figure 6-1 Steps and sources of error in ground-water
sampling.
Step
In-Situ Condition
Establishing a Sampling Point
Field Measurements
I
Sample Collection
Sample Delivery/Transfer
Field Blanks, Standards
Field Determinations
i
Preservation / Storage
Transportation
Sources of Error
Improper well construction/
placement; inappropriate
materials selection
Instrument malfunction;
operator error
Sampling mechanism bias;
operator error
Sampling mechanism bias;
sample exposure, degassing,
oxygenation; field conditions
Operator error; matrix
interferences
Instrument malfunction;
operator error; field conditions
Matrix interferences; handling/
labeling errors
Delay; sample loss
109
-------
There are two major obstacles to achieving control
over ground-water sampling errors. First, changes
which may occur in the integrity of samples prior to
sample delivery to the land surface cannot be
accounted for by the use of field blanks, standards
and split samples used in data quality assurance
programs. Second, most of the sources of error
which may affect sample integrity prior to sample
delivery are not well documented in the literature for
many of the contaminants of current interest. Among
these sources of error are the contamination of the
subsurface by drilling fluids, grouts or sealing
materials, the sorptive or leaching effects on water
samples due to well casing, pump or sampling tubing
materials' exposures and the effects on the solution
chemistry due to oxygenation, depressurization or gas
exchange caused by the sampling mechanism. These
sources of error have been investigated to some
extent for volatile organic contaminants under
laboratory conditions. However, to achieve confidence
in field monitoring and sampling instrumentation for
routine applications, common sense and a "research"
approach to regulatory monitoring may be needed.
Two of the most critical elements of a monitoring
program are establishing both reliable sampling points
and simple, efficient sampling protocols which will
yield data of known quality.
6.2 Establishing a Sampling Point
If adequate care is taken in the selection of drilling
methods, well construction materials and
development techniques, it should be possible to at
least approximate representative ground-water
sampling from a monitoring well. The representative
nature of the water samples can be maintained
consistently with a trained sampling staff and good
field-laboratory communication. Also, important
hydrologic measurements (i.e., water level, hydraulic
conductivity) can be made from the same sampling
point. A representative water sample may then be
defined as a minimally disturbed sample taken after
proper well purging which will allow the determination
of the chemical constituents of interest at
predetermined levels of accuracy and precision.
Sophisticated monitoring technology and sampling
instrumentation are poor substitutes for an
experienced sampling team which can follow a simple
proven sampling protocol.
This section details some of the considerations which
should be made in establishing a reliable sampling
point. There are a number of alternative approaches
to sampling point selection in monitoring network
design. Arrays of either nested monitoring wells or
multilevel devices (Barvenik and Cadwgan, 1983;
Pickens et a/., 1978) deployed at various sites within
the area of interest have their individual merits based
on the ease of verifying sampling point isolation,
durability, cost, ease of installation and site specific
factors. Deciding which option is most effective for
specific programs should be done with representative
sampling criteria in mind. The sampling points must
be durable, inert towards the chemical constituents of
interest, allow for purging of stagnant water, provide
sufficient water for analytical work with minimum
disturbance, and permit the evaluation of the
hydrologic characteristics of the formation of interest.
Monitoring wells can be constructed to meet these
criteria because a variety of drilling methods,
materials, sampling mechanisms and pumping
regimes for sampling and hydrologic measurements
can be selected to meet the current needs of most
monitoring programs.
The placement and number of wells will depend on
the complexity of the hydrologic setting and the
degree of spatial and temporal detail needed to meet
the goals of the program. It is important to note that
both the directions and approximate rates of ground-
water movement must be known in order to interpret
the chemical data satisfactorily. In this way it may
also be possible to estimate the nature and location
of pollutant sources (Gorelick ef a/., 1983).
Subsurface geophysical techniques can be very
helpful in determining the optimum placement of
monitoring wells under appropriate conditions and
when sufficient hydrogeologic information is available
(Evans and Schweitzer, 1984). Well placement should
be viewed as an evolutionary activity which may
expand or contract as the needs of the program
dictate.
6.2.7 Well Design and Construction
Effective monitoring well design and construction
requires considerable care and at least some
understanding of the hydrogeology and subsurface
geochemistry of the site. Preliminary borings, well
drilling experience and the details of the operational
history of a site can be very helpful. Monitoring well
design criteria include depth, screen size, gravel pack
specifications, and yield potential. These
considerations differ substantially from those applied
to production wells. The simplest, narrow diameter
well completions which will permit development,
accommodate the sampling gear and minimize the
need to purge large volumes of potentially
contaminated water are preferred for effective routine
monitoring activities. Helpful references are in several
publications (Barcelona ef a/., 1983; Scalf ef a/.,
1981; Wehrmann, 1983).
6.2.2 Well Drilling
The selection of a particular drilling technique should
depend on the geology of the site, the expected
depths of the wells and the suitability of drilling
equipment for the contaminants of interest. The
various drilling and well completion methods have
been reviewed with reference to these criteria in the
previous chapter. Regardless of the technique used,
every effort should be made to minimize subsurface
110
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disturbance. For critical applications, the drilling rig
and tools should be steam cleaned to minimize the
potential for cross-contamination between
formations or successive borings. The use of drilling
muds can be a liability for trace chemical constituent
investigations because foreign organic matter will be
introduced into the penetrated formations. Even
"clay" muds without polymeric additives contain some
organic matter which is added to stabilize the clay
suspension and may interfere with some analytical
determinations. Table 6-2 contains information on
the total and soluble organic carbon contents of some
common drilling and grouting materials (Barcelona,
1983). The effects of drilling muds on ground-water
solution chemistry have not been investigated in
detail. However, existing reports indicate that the
organic carbon introduced during drilling can cause
false water-quality observations for long periods of
time (Barcelona, 1984; Brobst, 1984). The fact that
the interferences are observable for gross indicators
of levels of organic carbon compounds (i.e., TOC)
and reduced substances (i.e., chemical oxygen
demand) strongly suggests that drilling aids are a
potential source of serious error. Innovative drilling
techniques may be called for in special situations
(Yare, 1975).
6.2.3 Well Development, Hydraulic Performance
and Purging Strategy
Once a well is completed, the sampling point must be
prepared for water sampling and measures must be
taken to evaluate its hydraulic characteristics. These
steps provide a basis for the maintenance of reliable
sampling points over the duration of a ground-water
monitoring program.
6.2.3.1 Well Development
The proper development of monitoring wells is
essential to the collection of "representative" water
samples. During the drilling process, fine particles are
forced through the sides of the bore hole into the
formation, forming a mud cake that reduces the
hydraulic conductivity of the materials in the
immediate area of the well bore. To allow water from
the formation being monitored to freely enter the
monitoring well, this mud cake must be broken down
opposite the screened portion of the well and the
fines removed from the well. This process also
enhances the yield potential of the monitoring well, a
critical factor when constructing monitoring wells in
low-yielding geologic materials.
More importantly, monitoring wells must be developed
to provide water free of suspended solids for
sampling. When sampling for metal ions and other
dissolved inorganic constituents, water samples must
be filtered and preserved at the well site at the time
of sample collection. Improperly developed monitoring
wells will produce samples containing suspended
sediments that may both bias the chemical analysis of
the collected samples and cause frequent clogging of
field filtering mechanisms. The additional time and
money spent for well development will expedite
sample filtration and result in samples that are more
representative of water chemistry in the formation
being monitored.
The development procedures used for monitoring
wells are similar to those used for production wells.
The first step in development involves the movement
of water at alternately high and low velocity into and
out of the well screen and gravel pack to break down
the mud cake on the well bore and loosen fine
Table 6-2 Composition of Selected Sealing and Drilling Muds
Ash
(%bywt)
Organic Content
(%bywt)
Soluble Carbon
(% bywt)
Soluble
Carbon in Total
Organic Content
(%bywt)
"Bentonite" muds/grouts
Volclay* ("-'90% montmorillonite)
Benseal'
"Organic" muds/drilling aids
Ez-Mudc (acrylamide-sodium acrylate copolymer
dispersed in food-grade oil
[normally used in 0.25% dilution])
Reverf (guar bean starch-based mixture)
98.2
88.5
17.5
1.6
1.8
11.5
21.5
98.4
<0.001
<0.001
17.9
33.8
94.4
3.7
2.1
85.6
°AII percentages determined on a moisture-free basis.
Trademark of American Colloid Co.
'Trademark of NL Baroid/NL Industries Inc.
•Trademark of Johnson Division, UOP Inc.
Source: Wood, 1976.
111
-------
particles in the borehole. This step is followed by
pumping to remove these materials from the well and
the immediate area outside the well screen. This
procedure should be continued until the water
pumped from the well is visually free of suspended
materials or sediments.
6.2.3.2 Hydraulic Performance of Monitoring
Wells
The importance of understanding the hydraulics of the
geologic materials at a site cannot be
overemphasized. Collection of accurate water level
data from properly located ar,d constructed wells
provides information on the direction of ground-
water flow (Chapter 4). The success of a monitoring
program also depends on knowledge of the rates of
travel of both the ground water and solutes. The
response of a monitoring well to pumping also must
be known to determine the proper rate and length of
time of pumping prior to collecting a water sample.
Hydraulic conductivity measurements provide a basis
for judging the hydraulic connection of the monitoring
well and adjacent screened formation to the
hydrogeologic setting. These measurements also
allow an experienced hydrologist to estimate an
optimal sampling frequency for the monitoring
program (Barcelona et a/., 1985).
Traditionally, hydraulic conductivity testing has been
conducted by collecting drill samples which were then
taken to the laboratory for testing. Several techniques
using laboratory permeameters are routinely used.
Falling head or constant head permeameter tests on
recompacted samples in fixed wall or triaxial test cells
are among the most common. The relative
applicability of these techniques is dependent on both
operator skill and methodology since calibration
standards are not available. The major problem with
laboratory test procedures is that one collects data on
recompacted geologic samples rather than geologic
materials under field conditions. Only limited work has
been done to date on performing laboratory tests on
"undisturbed" samples to improve the field
applicability of laboratory hydraulic conductivity
results.
Hydraulic conductivity is most effectively determined
under field conditions by pumping or slug testing. The
water level drawdown can be measured during
pumping. Alternatively, water levels are measured
after the static water level is depressed by application
of gas pressure or elevated by the introduction of a
slug of water. These procedures are rather straight-
forward for wells which have been properly
developed. The utility of these measurements is
obvious when one considers the extent of pumping
necessary to remove stagnant water from the
monitoring well and how much water must be
removed to establish a representative volume of
formation water for sampling.
Figure 6-2a Example of well purging requirement
estimating procedure (Barcelona et al., 1985).
Given:
48-foot deep, 2-inch diameter well
2-foot long screen
3-foot thick aquifer
static water level about 15 feet below land surface
hydraulic conductivity = 10~2 cm/sec
Assumptions:
A desired purge rate of 500 mL/min and sampling rate of 100
mL/min will be used.
Calculations:
One well volume = (48 ft -15 ft) x 613 mL/ft (2-inch diameter
well)
= 20.2 liters
Aquifer Transmissivity = hydraulic conductivity x aquifer thickness
= 10~" m/sec x 1 meter
= 10"" mVsec or 8.64 mVday
From Figure 6-2b:
At 5 minutes: 95% aquifer water and
(5minx0.5L/min)/20.2L
= 0.12 well volumes
At 10 minutes: 100% aquifer water and
(10minx0.5L/min)/20.2L
= 0.24 well volumes
It appears that a high percentage of aquifer water can be obtained
within a relatively short time of pumping at 500 mL*min''. This
pumping rate is below that used during well development to prevent
well damange or further development.
Figure 6-2b Percentage of aquifer water versus time for
different transmissitivities.
120 r-
100
-5 80
60
.
20
620.0 mVday
Transmissivhy 0.062 m2/day
Q = 500 mL/min
Diameter = 5.08 cm
10 15 20
Time (minutes)
25
30
112
-------
6.2.3.3 Well Purging Strategies
The number of well volumes to be pumped from a
monitoring well prior to the collection of a water
sample must be tailored to the hydraulic properties of
the geologic materials being monitored, the well
construction parameters, the desired pumping rate,
and the sampling methodology to be employed. There
is no one single number of well volumes to be
pumped that is best or fits all situations. The goal in
establishing a well purging strategy is to obtain water
from the geologic materials being monitored while
minimizing the disturbance of the regional flow
system and the collected sample. To accomplish this
goal a basic understanding of well hydraulics and the
effects of pumping on the quality of water samples is
essential. Water that has remained in the well casing
for extended periods of time (i.e., more than about
two hours) has had the opportunity to exchange
gases with the atmosphere and to interact with the
well casing material. The chemistry of water stored in
the well casing is unrepresentative of that in the
aquifer and it should not be collected for analysis.
Purge volumes and pumping rates should be
evaluated on a case by case basis.
Gibb (1981) has shown how the measurements of
hydraulic conductivity can be used to estimate the
well purging requirement. An example of this
procedure is shown in Figures 6-2a and 6-2b. In
practice, it may be necessary to test the hydraulic
conductivity of several wells within a network. The
calculated purging requirement should then be
verified by measurements of pH and specific
conductance during pumping to signal equilibration of
the water being collected.
The selection of purging rates and volumes of water
to be pumped prior to sample collection can also be
influenced by the anticipated water quality. In
hazardous environments where purged water must be
contained and disposed of in a permitted facility, it is
desirable to minimize the amount of purged water.
This can be accomplished by pumping the wells at
very low pumping rates (100 ml/min) to minimize the
drawdown in the well and maximize the percent
aquifer water delivered to the surface in the shortest
period of time. Pumping at low rates, in effect,
isolates the column of stagnant water in the well bore
and negates the need for its removal. This approach
is only valid in cases where the pump intake is placed
at the top of, or in, the well screen.
In summary, well purging strategies should be
established by (1) determining the hydraulic
performance of the well; (2) calculating reasonable
purging requirements, pumping rates, and volumes
based on hydraulic conductivity data, well
construction data, site hydrologic conditions, and
anticipated water quality; (3) measuring the well
purging parameters to verify chemical "equilibrated"
conditions; and (4) documenting the entire effort
(actual pumping rate, volumes pumped, and purging
parameter measurements before and after sample
collection).
6.2.3.4 Sampling Materials and Mechanisms
In many monitoring situations it is not possible to
predict the requirements which either materials for
well casings, pumps and tubing or pumping
mechanisms that the head and lift conditions must
meet in order to provide error-free samples of
ground water. Ideally these components of the
system should be durable and inert towards the
chemical properties of samples or the subsurface so
as to neither contaminate nor remove chemical
constituents from the water samples. Due to the long
duration of regulatory programs' requirements, well
casing materials in particular must be sufficiently
durable and nonreactive to last several decades. It is
generally much easier to substitute more appropriate
sampling pumps or pump/tubing materials as
knowledge of subsurface conditions improves than to
drill additional wells to replace inadequate well casing
or screen materials. Also there is no simple way to
account for errors which occur prior to handling a
sample at the land surface. Therefore, it is good
practice to carefully choose these components of the
sampling system which make up the rigid materials in
well casing/screens or pumps and the flexible
materials used in sample delivery tubing.
Rigid Materials. An experienced hydrologist can plan
well construction details based mainly on
hydrogeologic criteria, even for challenging situations
where a separate contaminant phase may be present
(Villaume, 1985). However, the question of the best
material for a specific monitoring application must be
addressed by considering subsurface geochemistry
and the likely contaminants of interest. Therefore,
strength, durability and inertness should be balanced
with cost considerations in the choice of rigid
materials for well casing, screens, pumps, etc.
Common well casing materials include TFE (Teflon),
PVC (polyvinyl chloride), stainless steel, and other
ferrous materials. The strength, durability and
potential for sorptive or leaching interferences on
chemical constituent determinations have been
reviewed in detail for these materials (Barcelona et
a/., 1985; Barcelona et a/., 1983). Unfortunately, there
is very little documentation of the severity or
magnitude of well casing interferences from actual
field investigations. This is the point at which
optimized monitoring network design takes on an
element of "research" or the need for systematic
evaluation of the components of the monitoring
installation.
Polymeric materials have the potential to absorb
dissolved chemical constituents and leach either
previously sorbed substances or components of the
polymer formulations. Similarly, ferrous materials may
adsorb dissolved chemical constituents and leach
113
-------
metal ions or corrosion products which may introduce
errors into the results of chemical analysis. This
potential in both cases is real, yet not completely
understood. The recommendations in the references
noted above can be summarized as follows:
o TeflonR is the well casing material least likely to
cause significant error in ground-water
monitoring programs focused on either organic or
inorganic chemical constituents. It has sufficient
strength for most applications at shallow depth
(i.e., <100 m) and is among the most inert
materials ever made. For deeper installations, it
can be linked to another material above the
highest seasonal, stagnant water level.
o Stainless steel (either 316 or 304 type) well
casing can be expected to, under noncorrosive
conditions, be the second least likely material to
cause significant error for organic chemical
constituent monitoring investigations. The release
of Fe, Mn or Cr may occur under corrosive
conditions. Organic constituent sorption effects
may also be significant sources of error after
corrosion processes have altered the virgin
surface.
o Rigid PVC well casing material with National
Sanitation Foundation approval should be used in
monitoring well applications when noncemented
or threaded joints are used and organic chemical
constituents are not expected to be of present or
future interest. Significant losses of strength,
durability and inertness (i.e., sorption or leaching)
may be expected under conditions where organic
contaminants are present in high concentration. It
should perform adequately in inorganic chemical
constituent studies when organic constituents are
not present in high concentrations and tin or
antimony species are not target chemical
constituents.
Monitoring wells made of appropriate materials and
screened over discrete sections of the saturated
thickness of geologic formations can yield a wealth of
chemical and hydrologic information. Whether or not
this level of performance is achieved may depend
frequently on the care taken in evaluating the
hydraulic performance of the sampling point.
Flexible Materials. Pump components and sample
delivery tubing may contact a water sample more
intimately than other components of a sampling
system, including storage vessels and well casing.
Similar considerations of inertness and
noncontaminating properties apply to tubing, bladder,
gasket and seal materials. Experimental evidence
(Barcelona et a/., 1985) has supported earlier
recommendations drawn from manufacturers'
specifications (Barcelona et a/., 1983). A summary is
provided in Table 6-3. Again, the care taken in
materials selection for the specific needs of the
sampling program can pay real dividends and
provides greater assurance of error-free sampling.
Sample Mechanisms. It is important to remember that
sampling mechanisms themselves are not protocols.
The sampling protocol for a particular monitoring
Table 6-3 Recommendations for Flexible Materials in Sampling Applications
Materials Recommendations
Polytetrafluoroethylene
(Teflon-)
Polypropylene
Polyethylene (linear)
PVC (flexible)
Viton-
Silicone (medical grade only)
Neoprene
Recommended for most monitoring work, particularly for detailed
organic analytical schemes. The material least likely to introduce
significant sampling bias or imprecision. The easiest material 10 clean
in order to prevent cross-contamination.
Strongly recommended for corrosive high dissolved solids solutions.
Less likely to introduce significant bias into analytical results than
polymer formulations (PVC) or other flexible materials with the
exception of Teflon".
Not recommended for detailed organic analytical schemes. Plasticizers
and stabilizers make up a sizable percentage of the material by weight
as long as it remains flexible. Documented interferences are likely with
several priority pollutant classes.
Flexible elastomeric materials for gaskets, 0-rings, bladder, and tubing
applications. Performance expected to be a function of exposure type
and the order of chemical resistance as shown. Recommended only
when a more suitable material is not available for the specific use.
Actual controlled exposure trials may be useful in assessing the
potential for analytical bias.
"Trademark of DuPont, Inc.
Source: Barcelona et al., 1981.
114
-------
network is basically a step-by-step written
description of the procedures used for well purging,
sample delivery to the surface and the handling of the
samples in the field. Once the protocol has been
developed and used in a particular investigation, it
provides a basis for modifying the program if the
extent or type of contamination requires more
intensive work. An appropriate sampling mechanism
is an important part of any protocol. Ideally, the
pumping mechanism should be capable of purging
the well of stagnant water at rates of liters or gallons
per minute and also delivering ground water to the
surface so that sample bottles may be filled at low
flow rates (i.e.,~100 ml/min-1 to minimize
turbulence and degassing of the sample. In this way
the criteria for representative sampling can be met
while keeping the purging and sample collection steps
simple. Nielsen & Yeates (1985) have reviewed the
types of sample collection mechanisms commercially
available (Anonymous, 1985) in line with the results of
research studies of their performance (Barcelona et
al., 1984; Stoltzenburg and Nichols, 1985). Examples
of types of pumps or other samplers are shown in
Figure 6-3 and they are described fully in a number
of references (Barcelona et al., 1985; Gillham et a/.,
1983; Scalf et a/., 1981). Given all of the varied
hydrogeologic settings and potential chemical
constituents of interest several types of pumps or
sampling mechanisms may be suitable for specific
applications. Figure 6-4 contains some
recommendations for reliable sampling mechanisms
given the sensitivity of the sample to error. The main
criteria for sampling pumps are the capabilities to
purge stagnant water from the well and deliver the
water samples to the surface with minimal loss of
sample integrity. Clearly, a mechanism that is proven
to provide accurate and precise samples for volatile
organic compound determinations should be suitable
for most chemical constituents of interest.
Now that a sampling point and the means to collect a
sample have been established, the next step is the
development of the detailed sampling protocol.
6.3 Elements of the Sampling Protocol
There are few aspects of this subject which generate
more controversy than the sampling steps which
make up the sampling protocol. Efforts to develop
reliable protocols and optimize sampling procedures
require particular attention to sampling mechanism
effects on the integrity of ground-water samples
(Barcelona et a/., 1984; Stolzenburg and Nichols,
1985), as well as the potential errors involved in well
purging, delivery tubing exposures (Barcelona et al.,
1985; Ho, 1983), sample handling and the impact of
sampling frequency on both the sensitivity and
reliability of chemical constituent monitoring results.
Quality assurance measures including field blanks,
standards, and split control samples cannot account
for errors in these steps of the sampling protocol
(Barcelona et al., 1985). Actually, the sampling
protocol is the focus of the overall study network
design (Nacht, 1983) and should be prepared flexibly
to be refined as the information on site conditions
improves.
Each step within the protocol has a bearing on the
quality and completeness of the information being
collected. This is perhaps best shown by the
progression of steps depicted in Figure 6-5.
Corresponding to each step is a goal and
recommendation for achieving the goal. The principal
utility of this description is that it provides an outlined
agenda for high quality chemical and water quality
data.
To insure maximum utility of the sampling effort and
resulting data, documentation of the sampling
protocol as performed in the field is essential. In
addition to noting the obvious information (i.e.,
persons conducting the sampling, equipment used,
weather conditions, and documentation of adherence
to the protocol and unusual observations) three basic
elements of the sampling protocol should be
recorded: (1) water level measurements made prior to
sampling, (2) the volume and rate at which water is
removed from the well prior to sample collection (well
purging), and (3) the actual sample collection
including measurement of well-purging parameters,
sample preservation, sample handling and chain of
custody.
6.3.7 Water-Level Measurement
Prior to the purging of a well or sample collection, it is
extremely important to measure and record the water
level in the well to be sampled. These measurements
are needed to estimate the amount of water to be
pumped from the well prior to sample collection. In
addition, this information can be useful when
interpreting monitoring results. Low water levels may
reflect the influence of a nearby production well. High
water levels compared to measurements made at
other times of the year are indicative of recent
recharge events. In relatively shallow monitoring
settings high water levels from recent natural
recharge events may result in dilution of the total
dissolved solids in the collected sample. Conversely,
if contaminants are temporarily held in an unsaturated
zone above the geologic zone being monitored,
recharge events may "flush" these contaminants into
the shallow ground water system and result in higher
levels of some constituents.
Documenting the nonpumping water levels for all
wells at a site will provide historical information on the
hydraulic conditions at the site. Analysis of this
information will reveal changes in flow paths and
serve as a check on the effectiveness of the wells to
monitor changing hydrologic conditions. It is very
useful to develop an understanding of the seasonal
115
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Figure 6-3 Schematic diagrams of common ground-water sampling devices (Neilsen and Yeates, 1985).
58:
Sample line
Lifting bail'
Discharge Check
Valve Assembly .
(Inside Body)
Perforated
Flow Tube
Bladder
Air Line
to Pressure
Water Flow
Bailer Line
Intake Check Valve
Assembly
(Inside Screen)
Annular
Space
I-—L
• Anti-Clogging
Screen
• 1-1/4"O.D. xTI.D.
Rigid Tubing,
Usually 18 to 36" Long
-Water Flow
Helical Rotor Electric
- 3/4" Diameter Ball Submersible Pump
1" Diameter Threaded Seat
5/16" Diameter Hole
Bailer
Cut-Away Diagram
of a Gas-Operated Bladder Pump
3/16" riser tube
112" gas drive tube
Compression tube fitting
Sampler body
Teflon seal
Porous filter
Gas Entry Tube
Sample Discharge Tube
Notes:
1. Sampler length can be increased
for special applications
2. Fabrication materials can be selected
to meet analysis requirements
and in situ chemical environment
3. Tubing sizes can be modified for
special applications
Polypropylene Tubing
Threaded Access Cap
- PVC Pipe
Check Valve
jrangement
- Slotted Well Screen
T „ _ Simple Slotted Well Point
Teflon Connector _ ;, . __, .. _. .
6mm ID
Gas-Drive Sampling Device
Glass Tubing
6mm OD
Tubing
Well Casing
Gas-Drive Sampler Designed
for Permanent Installation in a
Borehole (Barcad Systems)
Outlet
Peristaltic Pump
Sample Collection Bottle
116
-------
Figure 6-4 Matrix of sensitive chemical constituents and various sampling mechanisms.
Type of
constituent
Volatile
Organic
Compounds
Organometallics
Dissolved Gases
Well-purging
Parameters
Trace Inorganic
Metal Species
Reduced
Species
Major Cations
& Anions
Example of
constituent
Chloroform
TOX
CH3Hg
O2, CO2
pH, S-
Eh
Fe, Cu
N02-, S-
Na*, K*, Ca"
Mg*»
ci-, so4-
Positive
displacement
bladder pumps
Thief, in situ or
dual check valve
bailers
Mechanical
positive
displacement
pumps
INCREASING RELIABILITY OF SAMPLING MECHANISMS
NG SAMPLE SENSITIVITY +-
INCREAS
Superior
performance
for most
applications
Superior
performance
for most
applications
Superior
performance
for most
applications
Superior
performance
for most
applications
Maybe
adequate if well
purging is
assured
May be
adequate if well
purging is
assured
Maybe
adequate if well
purging is
assured
Adequate
Maybe
adequate if well
purging is
assured
May be ade-
quate if design
and operation are
controlled
May be ade-
quate if design
and operation are
controlled
Adequate
Adequate
Gas-drive
devices
Not recom-
mended
Not recom-
mended
Maybe
adequate
Adequate
Suction
mechanisms
Not recom-
mended
Not recom-
mended
May be ade-
quate if materials
are appropriate
Adequate
Figure 6-5 Generalized ground-water sampling protocol.
Step
Hydrologic Measurements
Well Purging
Sample Collection
Filtration/ Preservation
Field Determinations
Field Blanks/Standards
Sample Storage/Transport
Goal
Establish nonpumping water level.
Remove or isolate stagnant H20
which would otherwise bias repre-
sentative sample.
Collect samples at land surface
or in well-bore with minimal distur-
bance of sample chemistry.
Filtration permits determination of
soluble constituents and is a form of
preservation. It should be done in the
field as soon as possible after
collection.
Field analyses of samples will effec-
tively avoid bias in determining
parameters/constituents which do
not store well; e.g., gases, alkalinity,
pH.
These blanks and standards will
permit the correction of analytical
results for changes which may occur
after sample collection: preservation,
storage, and transport.
Refrigerate and protect samples to
minimize their chemical alteration
prior to analysis.
Recommendations
Measure the water level to ±0.3 cm (±0.01 ft).
Pump water until well purging parameters (e.g., pH,
T, Q-', Eh) stabilize to ± 10% over at least two
successive well volumes pumped.
Pumping rates should be limited to ~100 mL/min
for volatile organics and gas-sensitive parameters.
Filter: Trace metals, inorganic anions/cations,
alkalinity.
Do not filter: TOC, TOX, volatile organic com-
pound samples; other organic compound samples
only when required.
Samples for determining gases, alkalinity and
pH should be analyzed in the field if at all possible.
At least one blank and one standard for each
sensitive parameter should be made up in the field
on each day of sampling. Spiked samples are also
recommended for good QA/QC.
Observe maximum sample holding or storage periods
recommended by the Agency. Documentation of
actual holding periods should be carefully performed.
117
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changes in water levels and associated chemical
concentration variability at the monitored site.
6.3.2 Purging
The volume of stagnant water which should be
removed from the monitoring well should be
calculated from the analysis of field hydraulic
conductivity measurements. Rule-of-thumb
guidelines for the volume of water which should be
removed from a monitoring well prior to sample
collection can cause time delays and unnecessary
pumping of excess contaminated water. These rules
(i.e. three-, five- or 10-well volume) largely ignore
the hydraulic characteristics of individual wells and
geologic settings. One advantage of using the same
pump to both purge stagnant water and collect
samples is the ability to measure pH and specific
conductance fl-1 in an in-line flow cell. These
parameters aid in the verification of the purging
efficiency and also provide a consistent basis for
comparisons of samples from a single well or wells at
a particular site. Since pH is a standard variable for
aqueous solutions that is affected by degassing and
depressurization (i.e., loss of C02), in-line
measurements provide more accurate and precise
determinations than discrete samples collected by
grab sampling mechanisms.
The following example illustrates some of the other
advantages of verifying the purge requirement for
monitoring wells.
For example, the calculated well purging requirement
(e.g., >90 percent aquifer water) calls for the removal
of five well volumes prior to sample collection for a
particular well. Field measurements of the well
purging parameters have historically confirmed this
recommended procedure. During a subsequent
sampling effort, 12 well volumes were pumped before
stabilized well purging parameter readings were
obtained. Several possible causes could be explored:
(1) A limited plume of contaminants may have been
present at the well at the beginning of sampling and
inadvertently discarded while pumping in an attempt
to obtain stabilized indicator parameter readings; (2)
The hydraulic properties of the well have changed
due to silting or encrustation of the screen, indicating
the need for well rehabilitation or maintenance; (3)
The flow-through device used for measuring the
indicator parameters was malfunctioning; or (4) The
well may have been tampered with by the introduction
of a contaminant or relatively clean water source in
an attempt to bias the sample results.
Documentation of the actual well purging process
employed should be a part of a standard field
sampling protocol.
6.3.3 Sample Collection and Handling
Water samples should be collected when the solution
chemistry of the ground water being pumped has
stabilized as indicated by pH, Eh, Q-1 and T
readings.
In practice, stable sample chemistry is indicated when
the purging parameter measurements have stabilized
over two successive well volumes. First, samples for
volatile constituents, TOC, TOX and those
constituents which require field filtration or field
determination should be collected. Then large volume
samples for extractible organic compounds, total
metals or nutrient anion determinations should be
collected.
All samples should be collected as close as possible
to the well head. A "tee" fitting placed ahead of the
in-line device for measuring the well purging
parameters makes this more convenient. Regardless
of the sample mechanism in use or the components
of the sampling train, wells that are located upgradient
of a site, and therefore are expected to be
representative of background quality, should be
sampled first to minimize the potential for cross-
contamination followed by the wells that are located
downgradient of a site and may in fact contain
contaminants from the site. Laboratory detergent
solutions and distilled water should be used to clean
the sampling train between samples. An acid rinse
(0.1 N HCI) or solvent rinse (i.e., hexane or methanol)
may be used to supplement these cleaning steps if
necessary. Cleaning procedures should be followed
by distilled water rinses which may be saved to check
cleaning efficiency.
The order in which samples are taken for specific
types of chemical analyses should be decided by the
sensitivity of the samples to handling (i.e., most
sensitive first) and the need for specific information.
For example, the flow chart shown in Figure 6-6
depicts a priority order for a generalized sample
collection effort. The samples for organic chemical
constituent determinations are taken in decreasing
order in relation to sensitivity to handling errors, while
the inorganic chemical constituents, which may
require filtration, are taken afterwards.
There are instances which arise, even with properly
developed monitoring wells, that call for the filtration
of water samples. It should be evident, however, that
well development procedures which require two to
three hours of bailing, swabbing, pumping or air
purging at each well will save many hours of time in
sample filtration. Well development may have to be
repeated at periodic intervals to minimize the
collection of turbid samples. In this respect, it is
important to minimize the disturbance of fines which
accumulate in the well bore. This can be achieved by
careful placement of the sampling pump intake at the
top of the screened interval, low pumping rates, and
by avoiding the use of bailing techniques which
disturb sediment accumulations at the bottom of the
well.
118
-------
Figure 6-6 Generalized flow diagram of ground-water sampling steps (Barcelona et al., 1985).
STEP
Well Inspection
Well Purging
PROCEDURE
Hydrologic Measurements
Removal or Isolation of Stagnant Water
ESSENTIAL ELEMENTS
Water-Level
Measurements
Representative Water
Access
Sample Collection
Filtration*
Field
Determinations**
Preservation
Field Blanks
Standards
Determination of Well-Purging Parameters
(pH, Eh, T, Q-')»*
Unfiltered
Assorted Sensitive
Inorganic Species
NOr, NH4*, Fe(ll)
(as needed for good
QA/QC)
Field Filtered*
1
Volatile Organics, TOX
1
1
Dissolved
r
Gases, TOC
1
Large Volume Sam-
ples for Organic
Compound Determi-
nations
\
i
Alkalinity/ Acidity
1
Trace Metal Samples
S", Sensitive
Inorganics
Verification of
Representative Water
Sample Access
Sample Collection by
Appropriate Mechanism
Minimal Sample Handling
Head-Space
Free Samples
Minimal Aeration or
Depressurization
Minimal Air Contact,
Field Determination
Adequate Rinsing against
Contamination
Minimal Air Contact,
Preservation
Storage
Transport
Major Cations and
Anions
Minimal Loss of Sample
Integrity Prior to Analysis
* Denotes samples which should be filtered in order to determine dissolved constituents. Filtration should be accomplished preferably with in-
line filters and pump pressure or by N2 pressure methods. Samples for dissolved gases or volatile organics should not be filtered. In
instances where well development procedures do not allow for turbidity-free samples and may bias analytical results, split samples should
be spiked with standards before filtration. Both spiked samples and regular samples should be analyzed to determine recoveries from both
types of handling.
** Denotes analytical determinations which should be made in the field.
119
-------
It is advisable to refrain from filtering TOC, TOX or
other organic compound samples as the increased
handling required may result in the loss of chemical
constituents of interest. Allowing the samples to settle
prior to analysis followed by decanting the sample is
preferable to filtration in these instances. If filtration is
necessary for the determination of extractable organic
compounds, the filtration should be performed in the
laboratory by the application of nitrogen pressure. It
may be necessary to run parallel sets of filtered and
unfiltered samples with standards to establish the
recovery of hydrophobic compounds when samples
must be filtered. All of the materials' precautions used
in the construction of the sampling train should be
observed for filtration apparatus. Vacuum filtration of
ground-water samples is not recommended.
Water samples for dissolved inorganic chemical
constituents (e.g., metals, alkalinity and anionic
species) should be filtered in the field. The preferred
arrangement is an in-line filtration module which
utilizes sampling pump pressure for its operation.
These modules have tubing connectors on the inlet
and outlet parts and range in diameter from 2.5 to 15
cm. Large diameter filter holders, which can be
rapidly disassembled for filter pad replacement, are
the most convenient and efficient designs (Kennedy
et a/., 1976; Skougstad and Scarbo, 1968).
Representative sampling is the result of the execution
of a carefully planned sampling protocol which
establishes necessary hydrologic and chemical data
for each sample collection effort. An important
consideration for maintaining sample integrity after
collection is to minimize sample handling which may
bias subsequent determinations of chemical
constituents. Since opportunities to collect high
quality data for the characterization of site conditions
may be limited by time, it is prudent to conduct
sample collection as carefully as possible from the
beginning of the sampling period. It is preferable to
emphasize the need to risk error on the conservative
side when the doubt exists as to the sensitivity of
specific chemical constituents to sampling or handling
errors. Repeat sampling or analysis cannot make up
for lost data collection opportunities.
Samples collected for specific chemical constituents
may require modifications of recommended sample
handling and analysis procedures. Samples that
contain several chemicals and extended storage
periods can cause significant problems in this regard.
It is frequently more effective to perform a rapid field
determination of specific inorganic constituents (e.g.,
alkalinity, pH, ferrous iron, sulfide, nitrite or
ammonium) than to attempt sample preservation
followed by laboratory analysis of these samples.
Many samples can be held for the U.S. EPA
recommended maximum holding times after proper
preservation. These are shown in Table 6-4 which
has been modified slightly from Scalf et a/. (1981).
6.3.4 Quality Assurance/Quality Control
Planning for valid water quality data collection
depends upon both the knowledge of the system and
continued refinement of all sample handling or
collection procedures. As discussed in Section 6.2 of
this chapter, the need to begin QA/QC planning with
the installation of the sampling point cannot be over-
emphasized.
The use of field blanks, standards and spiked
samples for field QA/QC performance is analogous to
the use of laboratory blanks, standards and
procedural or validation standards. The fundamental
goal of field QC is to insure that the sample protocol
is being executed faithfully and that situations leading
to error are recognized before they seriously impact
the data. The use of field blanks and standards and
spiked samples can account for changes in samples
which occur during sample collection.
Field blanks and standards enable quantitative
correction for bias (i.e., systematic errors), which
arise due to handling, storage, transport and
laboratory procedures. Spiked samples and blind
controls provide the means to correct combined
sampling and analytical accuracy or recoveries for the
actual conditions to which the samples have been
exposed.
All QC measures should be performed for at least the
most sensitive chemical constituents for each
sampling date. Examples of sensitive constituents
would be benzene or trichloroethylene as volatile
organic compounds and lead or iron as metals. It is
difficult to use laboratory blanks alone for the
determination of the limits of detection or quantitation.
Laboratory distilled water may contain apparently
higher levels of volatile organic: compounds (e.g.,
methylene chloride) than those of uncontaminated
ground-water samples. The field blanks and spiked
samples should be used for this purpose, conserving
the results of lab blanks as checks on elevated
laboratory background levels.
Whether the ground water is contaminated with
interfering compounds or not, spiked samples provide
a basis for both the identification of the constituents
of interest and the correction of their recovery (or
accuracy) based on the recovery of the spiked
standard compounds. For example, if
trichloroethylene in a spiked sample is recovered at a
mean level of 80 percent (-20 percent bias), the
concentrations of trichloroethylene determined in the
samples for this sampling date may be corrected by a
factor of 1.2 for low recovery. Similarly, if 50 percent
recovery (-50 percent bias) is reported for the
spiked standard, it is likely that sample handling or
analytical procedures are out of control and corrective
measures should be taken at once. It is important to
120
-------
Table 6-4 Recommended Sample Handling and Preservation Procedures for a Detective Monitoring Program
Parameters
(Type)
Well Purging
pH (grab)
Q-1 (grab)
T (grab)
Eh (grab)
Contamination Indicators
pH, Q- (grab)
TOC
TOX
Water Quality
Dissolved gases
(02, CH4, C02)
Alkalinity /Acidity
(Fe, Mn, Na*,
K*, Ca",
Mg")
(P04-, Ch,
Silicate)
N03-
so4-
NH4*
Phenols
Volume
Required (ml)
1 Sample*
50
100
1000
1000
As above
40
500
10 mL minimum
100
Filtered under
pressure with
appropriate
media
All filtered
1000 mL
@50
100
50
400
500
Container
(Material)
T,S,P,G
T,S,P,G
T,S,P,G
T,S,P,G
As above
G,T
G,T
G,S
T,G,P
T,P
(T,P,G
glass only)
T,P,G
T,P,G
T,P,G
T,G
Preservation
Method
None; field det.
None; field det.
None; field det.
None; field det.
As above
Dark, 4°C
Dark, 4°C
Dark, 4°C
4°C/None
Field acidified
to pH <2 with
HNO3
4°C
4°C
4°C
4°C/H2SO4to
pH<2
4°C/H3P04to
pH<4
Maximum
Holding
Period
<1 hr**
<1 hr**
None
None
As above
24 hr
5 days
<24hr
<6 hr**/
<24hr
6 months***
24 hr/
7 days;
7 days
24 hr
7 days
24 hr/
7 days
24 hr
Drinking Water Suitability
As, Ba, Cd, Cr,
Pb, Hg, Se, Ag
F-
Remaining Organic
Parameters
Same as above
for water
quality cations
(Fe, Mn, etc.)
Same as chloride
above
Same as
above
Same as
above
Same as above
Same as above
As for TOX/TOC, except where analytical method calls for acidification
of sample
6 months
7 days
24 hr
*lt is assumed that at each site, for each sampling date, replicates, a field blank and standards must be taken at equal volume to those of
the samples.
"Temperature correction must be made for reliable reporting. Variations greater than ± 10% may result from longer holding period.
***ln the event that HN03 cannot be used because of shipping restrictions, the sample should be refrigerated to 4°C, shipped immediately,
and acidified on receipt at the laboratory. Container should be rinsed with 1:1 HN03 and included with sample.
Note: T = Teflon; S = stainless steel; P = PVC, polypropylene, polyethylene; G = borosilicate glass.
From Scalf etal., 1981.
121
-------
know if the laboratory has performed these
corrections or taken corrective action when they
report the results of analyses. It should be further
noted that many regulatory agencies require evidence
of QC and analytical performance but do not generally
accept data which has been corrected.
Field blanks, standards and blind control samples
provide independent checks on handling and storage
as well as the performance of the analytical
laboratory. It should be noted that ground-water
analytical data is incomplete unless the analytical
performance data (e.g., accuracy, precision,
detection, and quantitation limits) are reported along
with each set of results. Discussions of whether
significant changes in ground-water quality have
occurred must be tempered by the accuracy and
precision performance for specific chemical
constituents.
Table 6-5 is a useful guide to the preparation of field
standards, and spiking solutions for split samples. It is
important that the field blanks and standards be made
on the day of sampling and are subjected to all
conditions to which the samples are exposed. Field
spiked samples or blind controls should be prepared
by spiking with concentrated stock standards in an
appropriate background solution prior to the collection
of any actual samples. Additional precautions should
Table 6-5 Field Standard and Sample Spiking Solutions
be taken against the depressurization of samples
during air transport and the effects of undue exposure
to light during sample handling and storage. All of the
QC measures noted above will provide both a basis
for high quality data reporting and a known degree of
confidence in data interpretation. Well planned overall
quality control programs will also minimize the
uncertainty in long-term trends when different
personnel have been involved in sample collection
and analysis.
6.3.5 Sample Storage and Transport
The storage and transport of ground-water samples
are often the most neglected elements of the
sampling protocol. Due care must be taken in sample
collection, field determinations and handling.
Transport should be planned so as not to exceed
sample holding time before laboratory analysis. Every
effort should be made to inform the laboratory staff of
the approximate time of arrival so that the most
critical analytical determinations can be made within
recommended storage periods. This may require that
sampling schedules be adjusted so that the samples
arrive at the laboratory during working hours.
The documentation of actual sample storage and
treatment may be handled by the use of chain of
custody procedures. An example of a chain of
Stock Solution for Field Spike of Split Samples
Sample Type
Alkalinity
Anions
Volume
50 ml
1 L
Composition
Na*, HC03-
K*, Na*, CI-, SO,-
Field Standard
(Concentration)
10.0; 25 (ppm)
25, 50 (ppm)
Solvent
H20
H20
Concentration of
Components
10,000; 25,000 (ppm)
25,000; 50,000 (ppm)
Field Spike
Volume
(50 MU
(1 ml)
Cations
1 L
Trace Metals 1 L
TOC
TOX
Volatiles
40 mL
50 ml
40 ml
Extractables A 1 L
Extractables B 1 L
Extractables C 1 L
F-, NO3-, P04=, SI
Na*, K*
Ca**, Mg**, CI-, NCV
Cd**, Cu**, Pb**
Cr***, Ni2*, Ag*
Fe***, Mn**
Acetone
KHP
Chloroform
2,4,6 Trichlorophenol
Dichlorobutane, Toluene
Dibromopropane, Xylene
Phenol Standards
Polynuclear Aromatic
Standards
Standards as Required
5.0; 10.0 (ppm) H20, H* (acid) 5,000; 10,000 (ppm)
10.0; 25.0 (ppm) H2O, H* (acid) 10,000; 25,000 (ppm)
0.2; 0.5 (ppm-C)
1.8;4.5(ppm-C)
12.5; 25 (ppb)
12.5; 25 (ppb)
25; 50 (ppb)
25; 50 (ppb)
25; 50 (ppb)
25; 50 (ppb)
H20
H20/poly*
(ethylene glycol)
H20/poly*
(ethylene glycol)
Methanol**
Methanol
Methanol
200; 500 (ppm-C)
1,800; 4,500 (ppm-C)
12,500; 25 (ppm)
12,500; 25 (ppm)
25; 50 (ppm)
25; 50 (ppm)
25; 50 (ppm)
25; 50 (ppm)
(1 mL)
(1 mL)
(40 ^L)
(500 nL)
(40 MU
(1 mL)
(ImU
(1 mL)
*75:25 water/polyethylene glycol (400 amu) mixture.
"Glass distilled methanol.
Source: Barcelona et al., 1981.
122
-------
custody form is shown in Figure 6-7. Briefly, the
chain of custody record should contain the dates and
times of collection, receipt and completion of all the
analyses on a particular set of samples. It frequently
is the only record that exists of the actual storage
period prior to the reporting of analytical results. The
sampling staff members who initiate the chain of
custody should require that a copy of the form be
returned to them with the analytical report. Otherwise,
verification of sample storage and handling will be
incomplete.
Sample shipment arrangements should be planned to
insure that samples are neither lost nor damaged
enroute to the laboratory. There are several
commercial suppliers of sampling kits which permit
refrigeration by freezer packs and include proper
packing. It may be useful to include special labels or
distinctive storage vessels for acid-preserved
samples to accommodate shipping restrictions.
6.4 Summary
Ground-water sampling is conducted for a variety of
reasons ranging from detection or assessment of the
extent of a contaminant release to evaluations of
trends in regional water quality. Reliable sampling of
the subsurface is inherently more difficult than either
air or surface water sampling because of the
inevitable disturbances which well-drilling or
pumping can cause and the inaccessibility of the
sampling zone. Therefore, "representative" sampling
generally requires minimal disturbance of the
subsurface environment and the properties of a
representative sample are therefore scale dependent.
For any particular case, the applicable criteria should
be set at the beginning of the effort to judge
representativeness.
Reliable sampling protocols are based on the
hydrogeologic setting of the study site and the degree
of analytical detail required by the information needs
of the monitoring program. Quality control over water
quality data begins with the evaluation of the hydraulic
performance of the sampling point or well and the
proper selection of mechanisms and materials for well
purging and sample collection. All other elements of
the program and variables which effect data validity
which follow sample collection may be accounted for
by field blanks, standards and control samples.
Although research is needed on a host of topics
involved in ground-water sampling, defensible
sampling protocols can be developed to insure the
collection of data of known quality for many types of
programs. If properly planned and developed, long-
term sampling efforts can benefit from the
refinements which research progress will bring.
Careful documentation will provide the key to this
opportunity.
6.5 References
Anonymous. 1985. Monitoring Products: A Buyers
Guide. Ground Water Monitoring Review 5(3):33-45.
Barcelona, M.J., J.P. Gibb, J.A. Helfrich, and E.E.
Garske. 1985. Practical Guide for Ground-Water
Sampling. State Water Survey Contract Report 374,
U.S. Environmental Protection Agency, Robert S.
Kerr Environmental Research Laboratory, Ada, OK
and U.S. Environmental Protection Agency,
Environmental Monitoring and Support Laboratory,
Las Vegas, NV.
Barcelona, M.J., J.A. Helfrich, E. E. Garske, and J.P.
Gibb. 1984. A Laboratory Evaluation of Ground Water
Sampling Mechanisms. Ground Water Monitoring
Review 4(2):32-41.
Barcelona, M.J. 1984. TOC Determinations in Ground
Water. Ground Water 22(1): 18-24.
Barcelona, M.J., and E.E. Garske. 1983. Nitric Oxide
Interference in the Determination of Dissolved Oxygen
by the Azide-Modified Winkler Method. Analytical
Chemistry 55:965-967.
Barcelona, M.J., J.P. Gibb, and R.A. Miller. 1983. A
Guide to the Selection of Materials for Monitoring
Well Construction and Ground-Water Sampling.
Illinois State Water Survey Contract Report, USEPA-
RSKERL, EPA-600/S2-84-024. 78 pp.
Barcelona, M.J. 1983. Chemical Problems in
Ground-Water Monitoring. Proceedings of the Third
National Symposium on Aquifer Rehabilitation and
Ground Water Monitoring, May 24-27, 1983,
Columbus, OH.
Barcelona, M.J., J.A. Helfrich, and E.E. Garske.
1985. Sampling Tubing Effects on Ground Water
Samples. Analytical Chemistry 47(2):460-464.
Barvenik, M.J., and R.M. Cadwgan. 1983. Multi-
Level Gas-Drive Sampling of Deep Fractured Rock
Aquifers in Virginia. Ground Water Monitoring Review
3(4):34-40.
Brass, H.J., M.A. Feige, T. Halloran, J.W. Mellow, D.
Munch, and R.F. Thomas. 1977. The National
Organic Monitoring Survey: Samplings and Analyses
for Purgeable Organic Compounds. In: Drinking Water
Quality Enhancement through Source Protection,
edited by R.B. Pojasek. Ann Arbor Science
Publishers, Ann Arbor, Ml.
Brobst, R.B. 1984. Effects of Two Selected Drilling
Fluids on Ground Water Sample Chemistry.
Monitoring Wells, Their Place in the Water Well
Industry Educational Session, NWWA National
Meeting and Exposition, Las Vegas, NV.
Claasen, H.C. 1982. Guidelines and Techniques for
Obtaining Water Samples That Accurately Represent
123
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Figure 6-7 Sample chain of custody form.
CHAIN OF CUSTODY RECORD
Sampling Date Site Name
Well or Sampling Points:
Sample Seta for Each; Inorganic, Organic, Both
Inclusive Sample Numbers;
Company's Name Telephone ( )
Address
number street city state zip
Collector's Name Telephone ( )
Date Sampled Time Started Time Completed
Field Information (Precautions, Number of Samples, Number of Sample
Boxes, Etc.):
1.
name organization location
2.
name organization location
Chain of Possession (After samples are transported off-site or to
laboratory):
1. (IN)
signature title
(OUT)
name (printed) date/time of receipt
2. (IN)
signature title
__^ (OUT)
name (printed)date/time of receipt
Analysis Information;
Analysis Begun Analysis Completed
Aliquot (date/time) Initials (date/time) Initials
1.
2.
3.
U.
5.
124
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the Water Chemistry of an Aquifer. U.S. Geological
Survey Open File Report, Lakeland, CO.
Dunlap, W.J., J.F. McNabb, M.R. Scalf, and R.L.
Cosby. 1977. Sampling for Organic Chemicals and
Microorganisms in the Subsurface. U.S.
Environmental Protection Agency, Robert S. Kerr
Environmental Research Laboratory, Ada, OK.
Evans, R.B., and G.E. Schweitzer. 1984. Assessing
Hazardous Waste Problems. Environmental Science
and Technology 18(11):330A-339A.
Everett, L.G., and L.G. McMillion. 1985. Operational
Ranges for Suction Lysimeters. Ground Water
Monitoring Review 5(3):51-60.
Everett, L.G., L.G. Wilson, E.W. Haylman, and L.G.
McMillion. 1984. Constraints and Categories of
Vadose Zone Monitoring Devices. Ground Water
Monitoring Review 4(4).
Everett, L.G., L.G. Wilson, and L.G. McMillion. 1982.
Vadose Zone Monitoring Concepts for Hazardous
Waste Sites. Ground Water 20(3):312-324.
Fenn, D., E. Cocozza, J. Isbister, 0. Braids, B. Yare,
and P. Roux. 1977. Procedures Manual for Ground
Water Monitoring at Solid Waste Disposal Facilities.
EPA-530/SW611, U.S. Environmental Protection
Agency, Cincinnati, OH.
Gibb, J.P., R.M. Schuller, and R.A. Griffin. 1981.
Procedures for the Collection of Representative Water
Quality Data from Monitoring Wells. Illinois State
Water Survey Cooperative Report 7, Illinois State
Water Survey and Illinois State Geological Survey,
Champaign, IL.
Gillham, R.W., M.J.L. Robin, J.F. Barker, and J.A.
Cherry. 1983. Ground Water Monitoring and Sample
Bias. API Pub. 4367, American Petroleum Institute.
Grisak, G.E., R.E. Jackson, and J.F. Pickens. 1978.
Monitoring Gro undwater Quality: The Technical
Difficulties. Water Resources Bulletin 6:210-232.
Gorelick, S.M., B. Evans, and I. Remsan. 1983.
Identifying Sources of Ground Water Pollution: An
Optimization Approach. Water Resources Research
19(3):779-790.
Heaton, T.H.E., and J.C. Vogel. 1981. "Excess Air" in
Ground Water. Journal Hydrology 50:201-216.
Ho, J.S-Y. 1983. Effect of Sampling Variables on
Recovery of Volatile Organics in Water. Journal
American Water Works Association 12:583-586.
Kennedy, V.C., E.A. Jenne, and J.M. Burchard. 1976.
Backflushing Filters for Field Processing of Water
Samples Prior to Trace-Element Analysis. Open-
File Report 76-126. U.S.Geological Survey Water
Resources Investigations.
Lindau, C.W., and R. F. Spalding. 1984. Major
Procedural Discrepancies in Soil Extracted Nitrate
Levels and Nitrogen Isotopic Values. Ground Water
22(3):273-278.
Mackay, D.M., P.V. Roberts, and J.A. Cherry. 1985.
Transport of Organic Contaminants in Ground Water.
Environmental Science and Technology 19(5):384-
392.
Nacht, S.J. 1983. Monitoring Sampling Protocol
Considerations. Ground Water Monitoring Review
Summer:23-29.
National Council of the Paper Industry for Air and
Stream Improvement. 1982. A Guide to Groundwater
Sampling. Technical Bulletin 362, NCASI, New York,
NY.
Nielsen, D.M., and G.L. Yeates. 1985. A Comparison
of Sampling Mechanisms Available for Small-
Diameter Ground Water Monitoring Wells. Ground
Water Monitoring Review 5(2):83-99.
Pickens, J.F., J.A. Cherry, G.E. Grisak, W.F. Merritt,
and B.A. Risto. 1978. A Multilevel Device for
Ground-Water Sampling and Piezometric Monitoring.
Ground Water 16(5):322-327.
Robbins, G.A., and M.M. Gemmell. 1985. Factors
Requiring Resolution in Installing Vadose Zone
Monitoring Systems. Ground Water Monitoring
Review 5(3):75-80.
Scalf, M.R., J.F. McNabb, W.J. Dunlap, R.L. Cosby,
and J. Fryberger. 1981. Manual of Ground Water
Quality Sampling Procedures. National Water Well
Association, OH.
Schwarzenbach, R.P. et a/. 1985. Ground-Water
Contamination by Volatile Halogenated Alkanes:
Abiotic Formation of Volatile Sulfur Compounds Under
Anaerobic Conditions. Environmental Science and
Technology 19:322-327.
Sisk, S.W. 1981. NEIC Manual for
Groundwater/Subsurface Investigations at Hazardous
Waste Sites. U.S. Environmental Protection Agency,
Office of Enforcement, National Enforcement
Investigations Center, Denver, CO.
Skougstad, M.W., and G.F. Scarbo, Jr. 1968. Water
Sample Filtration Unit. Environmental Science and
Technology 2(4):298-301.
Stolzenburg, T.R., and D.G. Nichols. 1985.
Preliminary Results on Chemical Changes in Ground
Water Samples Due to Sampling Devices. Report to
Electric Power Research Institute, Palo Alto,
California, EA-4118. Residuals Management
Technology, Inc., Madison, Wl.
Tinlin, R.M., ed. 1976. Monitoring Groundwater
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U.S. Environmental Protection Agency, Environmental
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Todd, O.K., R.M. Tinlin, K.D. Schmidt, and L.G.
Everett. 1976. Monitoring Ground-Water Quality:
Monitoring Methodology. EPA-600/4-76-026, U.S.
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U.S. Geological Survey. 1977. National Handbook of
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U.S. Geological Survey, Office of Water Data
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Contaminated with Dense, Non-Aqueous Phase
Liquids (NAPLS). Ground Water Monitoring Review
5(2):60-74.
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166.
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126
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CHAPTER 7
GROUND-WATER TRACERS
The material presented in this chapter has been
condensed from the report Ground-Water Tracers
(Davis et a/., 1985).
7.1 General Characteristics of Tracers
As used in hydrogeology, a tracer is matter or energy
carried by ground water which will give information
concerning the direction of movement and/or velocity
of the water and potential contaminants which might
be transported by the water. If enough information is
collected, the study of tracers can also help with the
determination of hydraulic conductivity, porosity,
dispersivity, chemical distribution coefficients, and
other hydrogeologic parameters. A tracer can be
entirely natural, such as the heat carried by hot-
spring waters; it can be accidentally introduced, such
as fuel oil from a ruptured storage tank; or it can be
introduced intentionally, such as dyes placed in water
flowing within limestone caverns.
Understanding the potential chemical and physical
behavior of the tracer in ground water is the most
important criterion in selecting a tracer. A tracer
should travel with the same velocity and direction as
the water and not interact with solid material. For
most uses, a tracer should be nontoxic. It should be
relatively inexpensive to use and should be, for most
practical problems, easily detected with widely
available and simple technology. The tracer should be
present in concentrations well above background
concentrations of the same constituent in the natural
system which is being studied. Finally, the tracer itself
should not modify the hydraulic conductivity or other
properties of the medium being studied.
No one ideal tracer has been found. Because the
natural systems to be studied are so complex and the
requirements for the tracers themselves so
numerous, the selection and use of tracers is almost
as much an art as it is a science.
7.2 Public Health Considerations
Artificial introduction of tracers must be done with a
careful consideration of possible health implications.
Usually, investigations using artificially introduced
tracers must have the approval of local or State
health authorities, local citizens must be informed of
the tracer injections, and the results should be made
available to the public. Under some circumstances,
analytical work associated with tracer studies must be
done in appropriately certified laboratories.
7.3 Direction of Water Movement
To complete a tracer test using more than one well,
the general direction of ground-water movement
should be known. This is particularly true if the travel
of tracers is to be studied using two wells with ground
water flowing under a natural gradient.
Unfortunately, local flow directions may diverge widely
from directions predicted on the basis of widely
spaced water wells (Figure 7-1). It is not at all
uncommon to inject a tracer in a well and not be able
to intercept that tracer in another well just a few
meters away, particularly if the tracer flows under the
natural hydraulic gradient which is not disturbed by
pumping.
7.4 Travel Time
Travel time of a water particle can be estimated using
the equation:
where:
t = time taken by the average water particle to
move through distance AL
ne = effective porosity
K = hydraulic conductivity
Ah = hydraulic head drop.
If a tracer travels with the water, t is also the travel
time of the tracer. The use of this equation is
illustrated in Figure 7-2.
As can be seen in Equation 7-1, the expected travel
time for a given head drop is a function of the
distance squared (AL)2 and therefore increases very
rapidly with the distance, AL. Thus, a tracer test in
one region using a specific hydraulic head drop of Ah
over a distance of 1,000 m would take 10,000 times
127
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Figure 7-1 Divergence from predicted direction of
ground water.
Land Surface
Contoured on
Regional Data
-Water Table
Buried
Channel
Actual Movement Almost at Right
Angles to Direction Predicted by
Regional Water Levels
Figure 7-2 Example of water particle (and tracer) travel
time calculation.
Land Surface
r Aquifer
Aquifer
K
= 100 meters /day
If AL = 1000 meters
(.3) (1000)2
then t =
(100) (10)
= 300 days
as long as a test in another region over a distance of
10 m which has the same head drop, provided the
effective porosities and hydraulic: conductivities are
identical.
7.5 Sorption of Tracers and Related
Phenomena
Sorption occurs when a dissolved ion or molecule
becomes attached to the surface of a solid or
dissolves in the solid. The term "sorption," as used
here, includes the sum of the physical-chemical
phenomena of ion exchange, induced dipole
moments, hydrogen bonding, ligand exchange, and
chemical bonding. Two processes of sorption are
adsorption, a strictly surficial phenomenon, and
absorption, a phenomenon which involves movement
of material from solution to sites within the structure
of the solid phase. Most sorption processes
discussed here are relatively fast, reversible
reactions; that is, the dissolved constituent which is
sorbed from the water can be released to the water
again under favorable circumstances. Cation
exchange is probably the most familiar type of
adsorption, and is a good example of reversible
sorption.
Molecules of some tracers have a tendency to be
sorbed on the surfaces of solids for brief periods,
after which they move off the solid and into the water
again. If the water is moving, the tracer molecules
move at a slower rate than the water molecules,
because tracer molecules spend part of their time
sorbed on solids. Thus, the sorptive characteristics of
a tracer must be known in order to design meaningful
tracer experiments.
Certain tracers will be virtually unaffected by sorptive
processes. Those tracers are commonly called
"conservative" tracers because their concentrations,
and hence their direct relation to the moving ground
water, will be conserved if hydrodynamic dispersion is
not considered.
Although unlikely in most artificially introduced tracer
experiments, the possibility of mineral dissolution or
precipitation should always be kept in mind. As a
simple example, if the sulfate ion is used as a tracer
in water which moves through a natural bed of
gypsum, dissolution of the gypsum will undoubtedly
add sulfate to the ground water and may confuse the
interpretation of the experiment.
7.6 Hydrodynamic
Molecular Diffusion
Dispersion and
Two natural phenomena, hydrodynamic dispersion
and molecular diffusion, always work together to
dilute the concentrations of artificially injected tracers.
These phenomena are complex and their effects are
difficult to separate in field experiments. The two
phenomena are, however, theoretically quite distinct.
128
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Hydrodynamic dispersion is produced by natural
differences in the local ground-water velocities
related to the local differences in permeabilities
(Figure 7-3). Molecular diffusion is produced by
differences in chemical concentrations which tend to
be erased in time by the random motion of molecules
(Figure 7-4). Generally, short-term tracer
experiments in permeable material are affected
almost exclusively by hydrodynamic dispersion. In
contrast, the concentrations of natural tracers moving
very slowly in highly heterogeneous materials are
affected profoundly by molecular diffusion.
7.7 Practical Aspects
7.7.7 Planning a Test
The purpose and practical constraints of a potential
tracer test must be understood clearly prior to actual
planning of tracer tests. Is only the direction of water
flow to be determined? Are other parameters such as
travel time, porosity, and hydraulic conductivity of
interest? How much time is available for the test? If
answers must be obtained within a few weeks, then
tracer tests using only the natural hydraulic gradient
between two wells which are more than about 20
meters apart would normally be out of the question
because of the long time period needed for the tracer
to flow between the wells. Another primary
consideration is the budget. If several deep holes are
to be drilled, if packers are to be set to control
sampling or injection, and if hundreds of samples
must be analyzed in an EPA-certified laboratory,
then total costs could easily exceed $1 million. In
contrast, some short-term tracer tests may be
possible at costs of less than $1,000.
The initial step in determining the physical feasibility
of a tracer test is to collect as much hydrogeologic
information as possible concerning the field area. The
logs of the wells at the site to be tested, or logs of
the wells closest to the proposed site, should be
obtained. Logs will give some idea of the
homogeneity of the aquifer, layers present, fracture
patterns, porosity, and boundaries of the flow system.
Local or regional piezometric maps, or any published
reports on the hydrology of the area (including results
of aquifer tests) are valuable, as they may give an
indication of the hydraulic gradient and hydraulic
conductivity.
The hydrogeologic information is used to estimate the
direction and magnitude of the ground-water velocity
in the vicinity of the study area (Fetter, 1981). One
method to arrive at a local velocity estimate is the use
of water-level maps together with Darcy's Law if
transmissivity, aquifer thickness, and head values are
available. The second method involves using a
central well with satellite boreholes, and running a
preliminary tracer test. The classical method for
determining the regional flow direction is to drill three
boreholes at extremities of a triangle, with the sides
Figure 7-3 Variations in ground-water flow and
distribution of tracer due to hydrodynamic
dispersion.
General Direction of Water Motion
• Initial position
• Position after one hour
A Position after two hours
Initial Distribution Distribution of Tracer
of Tracer Particles _ Particles after One Hour
Distribution of Tracer
Particles after Two Hours
Distance
Figure 7-4 Movement by molecular diffusion.
Spot of Dye
Soaked Blotter—
(No Water Movement)
Initial Conditions
One Hour
Three Hours
One Day
129
-------
100 to 200 m apart (Figure 7-5). The water levels
are measured and the line of highest slope gives the
direction of flow. However, regional flow is generally
not as important as local flow in most tracer tests,
and the importance of having an accurate flow
direction cannot be overemphasized. Gaspar and
Oncescu (1972) described a method to determine
local flow direction by drilling five to six satellite wells
in the general direction of flow. The advantage of
knowing the general flow direction is that fewer
observation wells will eventually be drilled. If a
preliminary value of the magnitude of the natural
velocity of the aquifer is available, then the injection
or pumping rate necessary to obtain radial flow can
be determined. Also, when a velocity magnitude is
obtained from the preliminary test or available data, a
decision as to the distance from the injection well to
observation well(s) can be made. This decision
depends on whether the test is a natural flow or
induced flow (injection or pumping) type test. Natural
flow tests are less common due to the greater
amount of time involved.
Figure 7-5
Determining the direction of ground-water
flow.
Water Level
Elevation = 391 m
Observation Well #1
Water Level
Elevation = 387m
Observation Well #2
Area of
Proposed
Tracer Test
Water Level
Elevation = 382 m
Observation Well #3
Predicted Direction
of Ground-Water Flow
A second major consideration when planning a test is
which tracers are the best for the conditions at the
site and the objectives of the test. Samples should be
analyzed for background values of relevant
parameters, such as temperature, major ions, natural
fluorescence, fluorocarbons, etc. Choice of a tracer
will depend partially on which analytical techniques
are easily available and which background
constituents might interfere with these analyses.
Various analytical techniques incorporate different
interferences, and consultation with the chemist or
technician who will analyze the samples is necessary.
Determination of the amount of tracer to inject is
based on the natural background concentrations
detection limit for the tracer and the dilution expected.
If a value for porosity can be estimated, the volume of
voids in the medium can be calculated as the volume
of a cylinder with one well at the center and the other
a distance away. Adsorption, ion exchange, and
dispersion will decrease the amount of tracer arriving
at the observation well, but recovery is usually not
less than 20 percent (of the injected mass) for two-
hole tests using a forced recirculation system and
conservative tracers. The concentration should not be
increased so much that density effects become a
problem. Lenda and Zuber (1970) gave graphs which
can be used to estimate the approximate quantity of
tracer needed. The values are based on estimates of
the porosity and dispersion coefficient of the aquifer.
7.7.2 Types of Tracer Tests
The variety of tracer tests is almost infinite when one
considers the various combinations of tracer types,
local hydrologic conditions, injection methods,
sampling methods, and the geological setting of the
site.
Some of these varieties are shown in Figure 7-6.
The following sections discuss a few of the more
common types of tracer tests. Differences in the tests
are due to the parameters (such as velocity,
dispersion coefficient, and porosity) which are to be
determined, the scale of the test, and the number of
wells to be used.
7.7.2.1 Single-Well Techniques
Two techniques, injection/withdrawal and point
dilution, give values of parameters which are valid at
a local scale. Advantages of single-well techniques
are:
o Less tracer is required than for two-well tests
o The assumption of radial flow is generally valid,
so natural aquifer velocity can be ignored, making
solutions easier
o Knowledge of the exact direction of flow is not
necessary.
Injection/ Withdrawal. The single-well
injection/withdrawal (or pulse) technique results in a
value of pore velocity and the longitudinal dispersion
coefficient. The method assumes that porosity is
known or can be estimated with reasonable accuracy.
A given quantity of tracer is instantaneously added to
the borehole, the tracer is mixed, and then two to
three borehole volumes of fresh water are pumped in
to force the tracer to penetrate the aquifer. Only a
small quantity is injected so that natural flow is not
disturbed.
After a certain time, the borehole is pumped out at a
constant rate which is large enough to overcome the
natural ground-water flow. Tracer concentration is
measured with time or pumped volume. If
concentration is measured at various depths with
point samplers, relative permeability of layers can be
determined. The dispersion coefficient is obtained by
matching experimental breakthrough curves with
130
-------
Figure 7-6 Common configurations and uses for groundwater tracing.
Samo'ing Point
t
"•"•^ Fractured Rock
'.v.V.v Tracer
A. To determine if trash in sinkhole contributes to
contamination of spring.
Sampling Point
Cave
Stream
A Cave
1
^nm^^^^^^^
77777T7
B. To measure velocity of water in cave stream.
Sinking Stream
Sampling Point
Rise Stream
I Ris
C. To check source of water at rise in stream bed.
Sampling Point
D. To determine if tile drain from septic tank contributes to
contamination of well.
Three Different Tracers
Waste Water
vvasre vvaier ,»- i
Lagoon W Toilet | Landfill
^ '"" Ij^^^nmrTm^^ Sampling Point
E. To determine source of pollution from three possibilities.
^77
Water Table
Sampling Point
llilir
II1IU
mill
Mill
11(11
iniii
mm
F. To determine velocity and direction of ground-water flow under
natural conditions. Injection followed by sampling from same well.
nni
Water Table
»XY.
m
- Zone of Injected Tracer
G. To test precipitation of selected constituents on the aquifer material
by injecting multiple tracers into aquifer then pumping back the
injected water.
131
-------
Figure 7-6 (continued)
0
i
Sampling Point
/ / 1 / rj i
~—
<-j
T i i i i j i i i i
_Water_Jable
I "• •"*•*.
"."••*
— • — '~ "^-t-Tl.
e=D
' ' ' ' ' "< '
H. To test velocity of movement of dissolved material under
natural ground-water gradients.
Multi-Level Sampling
t t t
I. To test hydrodynamic dispersion in aquifer under natural
ground-water gradients.
Sample Point m
Pumped
Well
Injection
Well
•"-- Water
Table"
Sampling
Point
77T7TT7;
k Fractured
Granite
Packers
K. To determine the interconnect fractures between two uncased
holes. Packers are inflated with air and can be positioned as
desired in the holes.
Sampling Points
J. To test a number of aquifer parameters using a pair of wells
with forced circulation between wells.
L. To determine the direction and velocity of natural ground-water
flow by drilling an array of sampling wells around a tracer injection
well.
132
-------
Figure 7-6 (continued)
Sampling Point
at Pumping Well
M. To verify connection between surface water and well.
theoretical curves based on the general dispersion
equation. A finite difference method is used to
simulate the theoretical curves (Fried, 1975).
Fried concluded that the method is useful for local
information (2 to 4 m) and for detecting the most
permeable strata. An advantage of this test is that
nearly all of the tracer is removed from the aquifer at
the end of the test.
Borehole Dilution. This technique can be used to
measure the magnitude and direction of horizontal
tracer velocity and vertical flow.
The procedure is to introduce a known quantity of
tracer instantaneously into the borehole, mix it well,
and then measure the concentration decrease with
time. The tracer is generally introduced into an
isolated volume of the borehole using packers.
Radioactive tracers have been used for borehole
dilution tests, but other tracers can be used.
Some factors to keep in mind when conducting a
point dilution test are the homogeneity of the aquifer,
effects of drilling (mudcake, etc.), homogeneity of the
mixture of the tracer and the well water, degree of
tracer diffusion, and density effects.
The ideal condition for conducting the test is to use a
borehole with no screen or gravel pack. If a screen is
used, it should be next to the borehole as dead space
alters the results. Samples should be very small in
volume so that flow is not disturbed by its removal.
The direction of ground-water flow can be measured
in a single borehole by a method similar to point
dilution. A tracer (often radioactive) is introduced
slowly and without mixing. A section of the borehole
is usually isolated by packers. After some time, a
compartmental sampler (four to eight compartments)
within the borehole is opened. The direction of
minimum concentration corresponds to the flow
direction. A similar method is to introduce a
radioactive tracer and subsequently measure its
adsorption on the borehole or well screen walls by
means of a counting device in the hole. The method
is described in more detail in Gaspar and Oncescu
(1972).
7.7.2.2 Two-Well Techniques
There are two methods, one testing for uniform
(natural) flow and the other for radial flow. The
parameters measured (dispersion coefficient and
porosity) are assumed to be the same for both types
of flow.
Uniform Flow. A tracer is placed in one well without
disturbing the flow field and a signal is measured at
observation wells. This test can be used at a local (2
to 5 m) or intermediate (5 to 100 m) scale, but it
requires much more time than radial tests. The
direction and magnitude of the velocity must be
known quite precisely, or a large number of
observation wells are needed. The quantity of tracer
needed to cover a large distance can be expensive.
On a regional scale environmental tracers are
generally used, including seawater intrusion,
radionuclides, or stable isotopes of hydrogen and
oxygen. Manmade pollution has also been used. For
regional problems, a mathematical model is calibrated
with concentration versus time curves from field data,
and the same model is used to predict future
concentration distributions.
Analysis of local or intermediate scale uniform flow
problems can be done analytically, semianalytically, or
by curve-matching. Layers of different permeability
can cause distorted breakthrough curves, which can
usually be analyzed (Gaspar and Oncescu, 1972).
One- or two-dimensional models are available.
Analytical solutions can be found in Fried (1975) and
Lenda and Zuber (1970).
Radial Flow. These techniques are based on
imposing a velocity on the aquifer, and generally
solutions are easier if radial flow is much greater than
uniform flow. A value for natural ground-water
velocity is not obtained, but porosity and the
dispersion coefficient are obtained.
A diverging test involves constant injection of water
into an aquifer with a slug or continuous flow of tracer
introduced instantaneously into the injected water.
The tracer is detected at an observation well which is
not pumping. Very small samples are taken at the
observation well so that flow is not disturbed. Packers
can be used in the injection well to isolate an interval.
Sampling can be done with point samplers or an
integrated sample can be taken.
Converging tests involve introduction of the tracer at
an observation well, and another well is pumped.
Concentrations are monitored at the pumped well.
The tracer is often injected between two packers or
below one packer, and then two to three well bore
133
-------
volumes are injected to push the tracer out into the
aquifer. At the pumping well, intervals of interest are
isolated (particularly in fractured rock), or an
integrated sample is obtained.
A recirculating test
the pumped water
well. This tests a
formation because
360 degrees. The
canceling out the
Theoretical curves
(see Sauty, 1980).
is similar to a converging test, but
is injected back into the injection
significantly greater part of the
the wells inject to and pump from
flow lines are longer, partially
advantage of a higher gradient.
are available for recirculating tests
7.7.3 Design and Construction of Test Wells
In many tracer tests the construction of test wells is
the single most expensive part of the work. It also can
be the source of major difficulty if the construction is
not done properly.
Five common types of problems are encountered with
tracer tests. The first problem relates to site selection.
If heavy equipment is to be moved into an area, lack
of overhead clearance, narrow roads, poor bearing
capacities of bridges, and the lack of flat ground at
the site can all be major problems. Also, overhead
electrical power lines at the site should be avoided.
One of the most common hazards is accidental
grounding of power lines by drill rigs and auger stems
with subsequent electrocution of workers.
The second problem relates to the improper choice of
drilling equipment and the use of drilling fluids which
will affect the tracer tests. Certain drilling muds and
mud additives have a very high capacity for the
sorption of most types of tracers. The muds could
also clog small pores and alter the permeability of the
aquifer near the drill hole. The use of compressed air
for drilling may avoid some of these problems.
A third problem is the choice of casing diameter.
Ideally, packers should be used to isolate the zones
being sampled from the rest of the water in the well.
For a number of reasons which include economics,
insufficient time, and lack of technical training,
packers are often not used in tracer tests. In this
case, the diameter of the sampling well should be as
small as possible in order to minimize the amount of
"dead" water in the well during sampling. The
diameter, however, cannot be too small because the
well must be adequately cleaned after installation and
the well must accommodate bailers, pumps, and
other sampling equipment. Common casing diameters
used range from about 1 in to 4 in for relatively
shallow test holes to as much as 6 in to 8 in for very
deep tests.
The type of casing to be used is a fourth concern,
primarily if low-level concentrations of tracers are to
be used and particularly, if these tracers are organic
compounds or metallic cations. For plastic casings,
Teflon absorbs and releases less organics than does
PVC. Adhesives used to connect sections of plastic
pipes may be also a troublesome source of interfering
organic compounds. Metal casing could release trace
metals but it is generally superior to plastic casing in
terms of strength and sorptive characteristics.
Inexpensive metal casing, however, will have a short
life if ground waters are corrosive.
A fifth problem is the choice of filter construction for
the wells, which depends on the aquifer and the type
of test to be completed. If the aquifer being tested is
a very permeable coarse gravel and if the casing
diameter is small, then numerous holes drilled in the
solid casing may be adequate. In contrast, for a
single-well test with an alternating cycle of injection
and pumping of large volumes of water into and out
of loose, fine-grained sand, an expensive well
screen with a carefully placed gravel pack may be
required. Regardless of the type of filter used, it is
absolutely essential that the casing perforations,
gravel pack, or screen, as well as the aquifer at the
well, be cleaned of silt, clay, drilling mud, and other
material which would prevent the free movement of
water in and out of the well. This process of cleaning
or development is so critical that it should be
specified in clear terms in any contract related to well
construction.
7.7.4 Injection and Sample Collection
Injection equipment depends on the depth of the
borehole and the funds available. In very shallow
holes, the tracer can be lowered through a tube,
placed in an ampule which is lowered into the hole
and broken, or it can be just poured in. Mixing is
desirable and important for most types of tests and is
simple for very shallow holes. For example, a plunger
can be surged up and down in the hole or the release
of the tracer can be through a pipe with many
perforations. Flanges on the outer part of the pipe will
allow the tracer to be mixed by raising and lowering
the pipe. For deeper holes, tracers must be injected
under pressure and equipment can be quite
sophisticated. The equipment used in work conducted
in fractured rock by the Department of Hydrology at
the University of Arizona is described in detail in
Simpson et a/. (1983).
Sample collection can also be simple or
sophisticated. For tracing thermal pulses, only a
thermistor needs to be lowered into the ground water.
For chemical tracers at shallow depths, a hand pump
may be sufficient. Bailers can also be used, but they
mix the tracer in the borehole which, for some
purposes, should be avoided. A Teflon bottom-
loading bailer is described in Buss and Bandt (1981).
It may be desirable to clear the borehole before taking
a sample, in which case a gas-drive pump can be
used to evacuate the well. For a nonpumping system,
deciding how much water must be withdrawn from a
borehole in order to obtain a representative sample of
the water adjacent to the borehole is not a trivial
134
-------
Figure 7-7
Results of tracer tests at the Sand Ridge
State Forest, Illinois.
Injected 4/25
Injected 4/27
Amino G Acid
Rhodamine Wt
(Lissamine FF injected but
not detected)
Injection Well
I
10
1.0
ID
=
1 0.1
0.01
0.1
0.01
0.001
At 10 Feet
At 50 Feet
J
4/25 5/1 5/10 5/20
6/1 6/10
Date
7/1 7/10 7/20
problem. If not enough water is withdrawn, the
sample composition will be influenced by semistatic
water, which will normally fill much of the well. If too
much water is drawn, a gradient towards the well will
be created and the natural movement of the tracer
will be distorted. A common rule of thumb is to pump
out four times the volume of water in the well before
the sample is taken.
If existing wells which have been drilled for water-
supply purposes are used for tracer tests, extreme
care is required because of the complex relationship
among such variables as pumping rates, patterns of
water circulation within the well, and the yields of
different parts of the aquifers which are penetrated.
This complexity is usually reflected in the variability of
water chemistry as a well is being pumped (Keith ef
a/., 1982; Schmidt, 1977). Stated simply, for wells
drawing water from complex aquifers or a series of
aquifers, an analysis of a single water sample taken
at a given point in time cannot yield definitive
information about the water chemistry of any
individual zone.
The preservation and analysis of samples is covered
in Chapter 6 of this publication. Keith et a/. (1982)
also cover some of the practical problems involved
with sample collection, analyses, and quality control.
7.7.5 Interpretation of Results
The following remarks and figures are intended only
as a brief qualitative introduction to the interpretation
of the results of tracer tests. More extensive and
quantitative treatments are found in the works of such
authors as Halevy and Nir (1962), Theis (1963), Fried
(1975), Custodio (1976), Sauty (1978), Grisak and
Pickens (1980), and Gelhar (1982).
The basic plot of the concentration of a tracer as a
function of time or water volume passed through the
system is called a breakthrough curve. The
concentration is either plotted as the actual
concentration (Figure 7-7) or, quite commonly, as
the ratio of the measured tracer concentration at the
sampling point, C, to the input tracer concentration,
C0 (Figure 7-8).
The measured quantity which is fundamental for most
tracer tests is the first arrival time of the tracer as it
goes from an injection point to a sampling point. The
first arrival time conveys at least two bits of
information. First, it indicates that a connection for
ground-water flow actually exists between the two
points. For many tracer tests, particularly in karst
regions, this is all the information which is desired.
Second, an approximation of the maximum velocity of
ground-water flow between the two points may be
obtained if the tracer used is conservative.
Interpretations more elaborate than the two simple
ones mentioned depend very much on the type of
aquifer being tested, the velocity of ground-water
flow, the configuration of the tracer injection and
sampling systems, and the type of tracer or mixture
of tracers used in the test.
After the first arrival time, interest is most commonly
centered on the arrival time of the peak concentration
for a slug injection or, for a continuous feed of
tracers, the time since injection when the
concentration of the tracer changes most rapidly as a
function of time (Figure 7-9). In general, if
conservative tracers are used, this time is close to
the theoretical transit time of an average molecule of
ground water traveling between the two points.
If a tracer is being introduced continuously into a
ditch penetrating an aquifer, as shown in Figure 7-8,
135
-------
Figure 7-8 Tracer concentration at sampling well, C,
measured against tracer concentration at
input, Co-
Ditch Filled with
Tracer Having a
Concentration of C.
Sampling Well with
Water Having a
Tracer Concentration
of C
Time of First /
Arrival /
\
1 _ Timo nf MavitT
i ^y \
Rate of Chang
\
^Tracer Front
A. Tracer movement from injection ditch to sampling well.
1.0
o
<£ 0.5
o
0 01 i^~ 1 >• Time
A B
B. Breakthrough Curve.
Figure 7-9 Incomplete saturation of acquifer with tracer.
Ditch Filled with
Tracer Which
Supplies 1/4 of
Downgradient
Ground-Water
Flow- ' Sampling Well
A. Tracer does not fully saturate aquifer.
0.50
0.00
B. Breakthrough curve.
Time
then the ratio C/C0 will approach 1.0 after the tracer
starts to pass the sampling point. The ratio of 1.0 is
rarely approached in most tracer tests in the field,
however, because waters are mixed by dispersion
and diffusion in the aquifer and because wells used
for sampling will commonly intercept far more ground
water than has been tagged by tracers (Figure 7-9).
Ratios of C/Co in the range of between 10-5 and 2 x
10'1 are often reported from field tests.
If a tracer is introduced passively into an aquifer but
is recovered by pumping a separate sampling well,
then various mixtures of the tracer and the native
ground water will be recovered depending on the
amount of water pumped, the transmissivity of the
aquifer, the slope of the water table, and the shape of
the tracer plume. Keely (1984) has presented this
problem graphically with regards to the removal of
contaminated water from an aquifer.
With an introduction of a mixture of tracers, possible
interactions between the tracers and the solid part of
the aquifer may be studied. If interactions take place,
they can be detected by comparing breakthrough
curves of a conservative tracer with the curves of the
other tracers being tested (Figure 7-10). A common
strategy for these types of tracer tests is to inject and
subsequently remove the water containing mixed
tracers from a single well. If injection is rapid and
pumping to remove the tracer follows immediately,
then a recovery of almost all the injected conservative
tracer is possible. If the pumping is delayed, the
injected tracer will drift downgradient with the general
flow of the ground water and the percentage of the
recovery of the conservative tracer will be less as
time progresses. Successive tests using longer delay
times between injection and pumping can then be
used to estimate ground-water velocities in
permeable aquifers with moderately large hydraulic
gradients.
The methods of quantitative analyses of tracer
breakthrough curves are generally by curve-
matching of computer-generated type curves, or by
analytical methods.
7.8 Types of Tracers
7.8.7 Water Temperature
The temperature of water changes slowly as it
migrates through the subsurface because water has a
high specific heat capacity compared to most natural
materials. For example, temperature anomalies
associated with the spreading of warm wastewater in
the Hanford Reservation in south central Washington
have been detected more than 8 km (5 mi) from the
source (U.S. Research and Development
Administration, 1975).
Water temperature is a potentially useful tracer,
although it has not been used frequently. The method
should be applicable in granular media, fractured
136
-------
Figure 7-10 Breakthrough curves for conservative and
nonconservative tracers.
o
o
0.10
0.05
0.00
Tracer A
(Conservative)
Tracer D
(Precipitated)
Tracer B
(Some Sorption) Trgcer c
(Largely Sorbed)
Time
Figure 7-11 Results of field test using a hot water tracer.
27.0
G 26.0
O
o> 25.0
| 24.0
I 23.0
22.0
21.0
6
* Well 1
• Well 2
. Well 3
o Well 4
Initial Temperature of Injected Fluid = 47.1 °C
30 50 70 90 110 130
Time After Injection (Minutes)
150
rock, or karst regions. Keys and Brown (1978) traced
thermal pulses resulting from the artificial recharge of
playa lake water into the Ogallala Formation in Texas.
They described the use of temperature logs
(temperature measurements at intervals in cased
holes) as a means of detecting hydraulic conductivity
differences in an aquifer. Temperature logs have also
been used to determine vertical movement of water in
a borehole (Keys and MacCary, 1971; Sorey, 1971).
Changes in water temperature are accompanied by
changes in density and viscosity of the water. This in
turn alters the velocity and direction of flow of the
water. For example, injected ground water with a
temperature of 40 °C will travel more than twice as
fast in the same aquifer under the same hydraulic
gradient as water at 5°C. Because the warm water
has a slightly lower density than cold water, buoyant
forces give rise to flow which "floats" on top of the
cold water. In order to minimize problems of
temperature-induced convection, small temperature
differences with very accurate temperature
measurements should be used if hot or cold water is
in the introduced tracer.
Temperature was used as a tracer for small-scale
field tests, using shallow drive-point wells two feet
apart in an alluvial aquifer. The transit time of the
peak temperature was about 107 min, while the
resistivity data indicated a travel time of about 120
min (Figure 7-11). The injected water had a
temperature of 38°C, while the ground-water
temperature was 20 °C. The peak temperature
obtained in the observation well was 27°C.
In these tests, temperature served as an indicator of
breakthrough of the chemical tracers, aiding in the
timing of sampling. It was also useful as a simple,
inexpensive tracer for determining the correct
placement of sampling wells.
Another application of water-temperature tracing is
the detection of river recharge in an aquifer. Most
rivers have large seasonal water temperature
fluctuations. If the river is recharging an aquifer, the
seasonal fluctuations can be detected in the ground
water adjacent to the river (Rorabaugh, 1956).
7.8.2 So//cf Particles
Solid material in suspension can be a useful tracer in
areas where water flows in large conduits such as
some basalt, limestone, or dolomite aquifers. Aley
(1976) reported that geese, bales of hay, and wheat
chaff have been used in Missouri in karst regions. In
the past decade, small particulate tracers such as
bacteria have been used successfully in porous
media.
Paper and Simple Floats. Some examples of these
tracers are small bits of paper (as punched out from
computer cards, for example), or multicolored
polypropylene floats. Due to the large size of these
tracers, they are useful only when flow is through
large passages. The particles must be of such a size
and density as to pass through shallow sections of
flow without settling out. Because these particulates
generally float on the surface, they travel faster than
the water's mean velocity. These tracers are most
137
-------
flow velocity and
useful for approximating the
establishing the flow path.
Dunn (1963) described the use of polypropylene
floats of approximately 3/32-in diameter and 1-in
length.
Signal-Emitting Floats. These are delayed time
bombs which float through a cave system. When the
bomb explodes, the location of the explosion is
determined by seismic methods at the surface
(Arandjelovic, 1969 and 1977). Problems with this
method include noise interference from wind, traffic,
and surface streams. Because these methods are
relatively expensive, they have seldom been used.
Yeast. The use of baker's yeast (Saccharomyces
cerevisiae) as a ground-water tracer in a sand and
gravel aquifer has been reported by Wood and Ehrlich
(1978). Yeast is a single-celled fungus which is
ovoid in shape. The diameter of a yeast cell is 2 to 3
lam, which closely approximates the size of
pathogenic bacterial cells. This tracer is probably
most applicable in providing information concerning
the potential movement of bacteria.
Wood and Ehrlich (1976) found that the yeast
penetrated more than 7 m into a sand and gravel
aquifer in less than 48 hours after injection. The
tracer is very inexpensive, as is analysis. The lack of
environmental concerns related to this tracer is
another of its advantages.
Bacteria. Bacteria are the most commonly used
microbial tracers, due to their ease of growth and
simple detection. Keswick et al. (1982) reviewed case
studies of bacteria used as tracers. Some of the
bacteria which have been used successfully are
Escherichia coliform (E. coli), Streptococcus faecalis,
Bacillus stearothermophilus, Serratia marcescens,
and Serratia indica. They range in size from 1 to 10
m and have been used in a variety of applications.
A fecal coliform, E. coli, has been used to indicate
fecal pollution at pit latrines, septic fields, and sewage
disposal sites. A "marker" such as antibiotic
resistance or H2S production is necessary to
distinguish the tracer from background organisms.
The greatest health concern in using these tracers is
that the bacteria must be nonpathogenic to man.
Even E. coli has strains which can be pathogenic,
and Davis et al. (1970) reported that Serratia
marcescens may be pathogenic. Antibiotic-resistant
strains are another concern. The antibiotic resistance
can be transferred to potential human pathogens.
This can be avoided by using bacteria which cannot
transfer this genetic information. As is true with most
other injected tracers, permission to use bacterial
tracers should be obtained from the proper Federal,
State, and local health authorities.
Viruses. Animal, plant, and bacterial viruses have
been used as ground-water tracers. Viruses are
generally much smaller than bacteria, ranging from
0.2 to 1.0 urn (see Table 7-1). In general, human
enteric viruses cannot be used due to disease
potential. Certain vaccine strains, such as a type of
polio virus, have been used but are considered risky.
Most animal enteric viruses are considered safer as
they are not known to infect man (Keswick et a/.,
1982). However, neither human nor most animal
viruses are generally considered to be suitable tracers
for field work because of their potential to infect man.
Table 7-1 Comparison of Microbial Tracers
Size
Tracer
Time
Required for
Assay (days)
Essential
Equipment
Required
Bacteria
Spores
Yeast
Viruses:
Animal (enteric)
Bacterial
1-10
25-33
2-3
0.2-0.8
0.2-1.0
1-2
1/2
12
3-5
1/2-1
Incubator*
Microscope
Plankton nets
Incubator*
Incubator
Tissue culture
Laboratory
Incubator*
*Many may be assayed at room temperature.
Spores. Lycopodium spores have been used as a
water tracer since the early 1950s, and the
techniques are well developed. Spore tracing was
initiated by Mayr (1953) and Maurin and Zotl (1959)
and modified by Drew (1968). Lycopodium is a
clubmoss which has spores that are nearly spherical
in shape, with a mean diameter of 33 um. It is
composed of cellulose and is slightly denser than
water, requiring some turbulence to keep the material
in suspension. Some advantages of lycopodium are:
o The spores are relatively small
o They are not affected by water chemistry or
adsorbed by clay or silt
o They travel at approximately the velocity of the
surrounding water
o The injection concentration can be very high
(e.g., 8 x 106 spores per cm3)
o They pose no health threat
o The spores are easily detectable under the
microscope
o At least five dye colors may be used, allowing five
tracings to be conducted simultaneously in a karst
system.
138
-------
Some disadvantages associated with its use include
the large amount of time required for preparation and
analysis of the spores, and the problem of spores
being filtered by sand or gravel if flow is not
sufficiently turbulent.
The basic procedure involves the addition of a few
kilograms of dyed spores to a cave or sinking stream.
The movement of the tracer is monitored by sampling
downstream in the cave or at a spring, with plankton
nets installed in the stream bed. The sediment caught
in the net is concentrated and treated to remove
organic matter. The spores are then examined under
the microscope.
Tracing by lycopodium spores is most useful in open
joints or solution channels (karst terrain). It is not
useful in wells or boreholes unless the water is
pumped continuously to the surface and filtered. A
velocity of a few miles per hour has been found
sufficient to keep the spores in suspension.
According to Smart and Smith (1976), lycopodium is
preferable to dyes for use in large-scale water
resource reconnaissance studies in karst areas. This
holds if skilled personnel are available to sample and
analyze the spores and a relatively small number of
sampling sites are used.
The spores survive well in polluted water, but do not
perform well in slow flow or in water with a high
sediment concentration. Lycopodium spores have
been used extensively in the United States, Great
Britain, and other countries to determine flow paths
and to estimate time of travel in karst systems.
7.8.3 Ions
Ionic compounds such as common salts have been
used extensively as ground-water tracers. This
category of tracers includes those compounds which
undergo ionization in water resulting in separation into
charged species possessing a positive charge
(cations) or a negative charge (anions). The charge
on an ion affects its movement through aquifers by
numerous mechanisms.
Ionic tracers have been used as tools for a wide
range of hydrologic problems dealing with the
determination of flow paths and residence time and
the measurement of aquifer properties.
Specific characteristics of individual ions or ionic
groups may approach those of an ideal tracer,
particularly in the case of dilute concentrations of
certain anions.
In most situations, anions (negatively charged ions)
are not affected by the aquifer medium. Mattson
(1929), however, has shown that the capacity of clay
minerals for holding anions increases with decreasing
pH. Under conditions of low pH, anions in the
presence of clay, other minerals, or organic detritus
may undergo anion exchange. Other effects which
may occur include anion exclusion and
precipitation/dissolution reactions. Cations (positively
charged ions) react much more frequently with clay
minerals through the process of cation exchange
which in turn displaces other cations such as sodium
and calcium into solution. For this reason, little work
has been done with cations due to the interaction with
the aquifer media. Kaufman (1956) has shown that
when permeabilities and flow rates are low, often
indicative of a large clay fraction, the solid phase may
have a considerable adsorption of an ionic
component. This is significant for cationic tracers and
may have some significance for certain anionic
tracers.
One advantage of the simple ionic tracers is that they
do not decompose and thus are not lost from the
system. However, a large number of ions (including
CI" and NOa") have high natural background
concentrations, thus requiring the injection of a tracer
of high concentration. This may result in density
separation and gravity segregation during the tracer
test (Grisak, 1979). Density differences will alter flow
patterns, the degree of ion exchange, and secondary
chemical precipitation, which may change the aquifer
permeability.
Various applications of ionic tracers have been
described in the literature. Methods similar to those
used for CI" were also postulated for ions such as
nitrate (NOs"), dichromate (CF^O/), and ammonium
(NH4 + ) (Haas, 1959). Murray (1981) used lithium
bromide (LiBr) in carbonate terraine to establish
hydraulic connection between a landfill and a fresh-
water spring where use of rhodamine WT dye tracer
proved inappropriate. Sodium chloride (NaCI) was
used by Mather (1969) to investigate the influence of
mining subsidence on the pattern of ground-water
flow. Tennyson (1980) used bromide (Br") to
evaluate pathways and transit time of recharge
through soil at a proposed sewage effluent irrigation
site. Chloride (CI") and calcium (Ca + ) were used by
Grisak (1979) to study solute transport mechanisms
in fractures. Potassium (K + ) was used to determine
leachate migration and extent of dilution by receiving
waters located by a waste disposal site (Ellis, 1980).
7.8.4 Dyes
Dyes are relatively inexpensive, simple to use, and
effective. Nonfluorescent dyes include congo red and
malachite green, which have been used in
conjunction with cotton strip detectors (Drew, 1968)
or with visual detection, often in soil studies. The
extensive use of fluorescent dyes for ground-water
tracing began around 1960. Fluorescent dyes are
preferable to nonfluorescent varieties due to much
better detectability.
Although fluorescent dyes exhibit many of the
properties of an ideal tracer, a number of factors
interfere with concentration measurement.
Fluorescence is used to measure dye concentration,
139
-------
but it may vary with suspended sediment load,
temperature, pH, CaCOs content, salinity, etc. Other
variables which affect tracer test results are
"quenching" (some emitted fluorescent light is
reabsorbed by other molecules), adsorption, and
photochemical and biological decay. Another
disadvantage of fluorescent dyes is their poor
performance in tropical climates due to chemical
reactions with dissolved carbon dioxide.
The advantages of using these dyes include their
very high detectability, rapid field analysis, and
relatively low cost and low toxicity.
Fluorescence intensity is inversely proportional to
temperature. Smart and Laidlaw (1977) described the
numerical relationship and provided temperature
correction curves. The effect of pH on rhodamine WT
fluorescence is shown in Figure 7-12. An increase in
the suspended sediment concentration generally
causes a decrease in fluorescence. Adsorption on
kaolinite caused a decrease in the measured
fluorescence of several dyes, as measured by Smart
and Laidlaw.
The detected fluorescence may decrease or actually
increase due to adsorption. If dye is adsorbed onto
suspended solids, and the fluorescence
measurements are taken without separating the water
samples from the sediment, the dye concentration is
a measurement of sediment content and not of water
flow. Adsorption can occur on organic matter, clays
(bentonite, kaolinite, etc.), sandstone, limestone,
plants, plankton, and even glass sample bottles.
These adsorption effects are a strong incentive to
choose a nonsorptive dye for the type of medium
tested. The sorption of orange dyes on bentonite clay
is shown in Table 7-2.
Dyes travel slower than water due to adsorption, and
are generally not as conservative as the ionic or
radioactive tracers. Drew (1968) compared
lycopodium, temperature, and fluorescein as karst
tracers and found fluorescein breakthrough to be the
slowest (Figure 7-13).
Although only one test is generally run due to
economic considerations, it may be advisable to run
several tests to check reproducibility if accuracy is
important. Brown and Ford (1971) obtained some
very interesting results by running three identical dye
tracer tests in the same karst system. These yielded
three different flow-through times. One of the values
differed by 50 percent from the original test value.
Fluorescein, also known as uranin, sodium
fluorescein, and pthalien, has been one of the most
widely used green dyes. Like all green dyes, its use
is commonly complicated by high natural background
fluroescence, which lowers sensitivity of analyses and
makes interpretation of results more difficult.
Feuerstein and Selleck (1963) recommended that
Table 7-2 Measured Sorption of Dyes on Bentonite Clay
Dye
Rhodamine WT
Rhodamine B
Sulfo Rhodamine B
Losses Due to
Adsorption on Clay
28
96
65
Source: Repogle et al., 166.
Figure 7-12 The effect of pH on rhodamine WT (adapted
from Smart and Laidlaw, 1977).
HCI &NaOH
HN03 & NaOH
3.0
5.0
7.0
9.0
11.0
pH
Figure 7-13 A comparison of the results of three
simultaneous tracer tests in a karst system
(data from Drew, 1968).
=5 6
D
C _«
•~ ^ 5
III 4
S o 8 3
p- p c
E ^ Q „
Q. "<
O
Lycopodium
Temperature
._«_ Fluorescein
5.0
10.0 15.0
Time (minutes)
20.0
25.0
140
-------
Table 7-3 Sensitivity and Minimum Detectable Concentrations for the Tracer Dyes
Dye
Amino G Acid
Photine CU
Fluorescein
Lissamine FF
Pyranine
Rhodamine B
Rhodamine WT
Sulfo Rhodamine B
Sensitivity*
^g/L Per Scale Unit
0.27
0.19
0.11
0.11
0.333
0.010
0.013
0.061
Background
Reading**
Scale Units
0-100
19.0
19.0
26.5
26.5
26.5
1.5
1.5
1.5
Minimum
Detectability***
MI/L
0.51
0.36
0.29
0.29
0.087
0.010
0.013
0.061
For a Turner 111 filter fluorometer with high-sensitivity door and recommended filters and lamp at 21 °C.
* At a pH of 7.5.
** For distilled water.
*** For a 10 percent increase over background reading or one scale unit, whichever is larger.
Adapted from Smart and Laidlaw, 1977.
fluorescein be restricted to short-term studies of
only the highest quality water.
Lewis et al. (1966) used fluorescein in a fractured
rock study. Another example is a mining subsidence
investigation in South Wales, where more than one
ton of fluorescein was used in a sandstone tracer test
(Mather et al., 1969). Tester et al. (1982) used
fluorescein to determine fracture volumes and
diagnose flow behavior in a fractured granitic
geothermal reservoir. He found no measurable
adsorption or decomposition of the dye during the 24
hr exposures to rocks at 392 °F. Omoti and Wild
(1979) stated that fluorescein is one of the best
tracers for soil studies, but Rahe et al. (1978) did not
recover any injected dye in their hillslope studies,
even at a distance of 2.5 m downslope from the
injection point. The same experiment used bacterial
tracers successfully. Figure 7-13 compares
fluorescein, lycopodium, and temperature as karst
tracers.
The approximate sensitivity and minimum detection
limit for a number of dyes are given in Table 7-3.
Another green fluorescent dye, pyranine, has been
used in several soil studies, and Reynolds (1966)
found it to be the most stable dye used in an acidic,
sandy soil. Omoti and Wild (1979) recommended
pyranine and fluorescein as the best tracers for soil
tests, although pyranine is relatively unstable if the
organic matter content of the soil is high. Drew and
Smith (1969) stated that pyranine is not as easily
detectable as fluorescein, but is more resistant to
decoloration and adsorption. Pyranine has a very high
photochemical decay rate, and is strongly affected by
pH in the range found in most natural waters
(Mclaughlin, 1982).
The orange dye rhodamine WT, is thought to be
slightly less toxic than rhodamine B and sulfo
rhodamine B (Smart and Laidlaw, 1977). This source
notes that rhodamine WT and fluorescein are of
comparable toxicity, but Aley and Fletcher (1976)
stated that rhodamine WT is not as "biologically safe"
as fluorescein.
This dye has been considered one of the most useful
tracers for quantitative studies, based on minimum
detectability, photochemical and biological decay
rates, and adsorption (Smart and Laidlaw, 1977;
Wilson, 1968; and Knuttson, 1968). Hubbard et al.
(1982) stated that it is the most conservative of dyes
available for stream or karst tracing.
Some recent uses of rhodamine WT include projects
by Burden (1981), Aulenbach et al. (1978), Brown
and Ford (1971), Gann (1975), and Aulenbach and
Clesceri (1980). Burden successfully used the dye in
a water contamination study in New Zealand in an
alluvial aquifer. Aulenbach and Clesceri also found
rhodamine WT very successful in a sandy medium.
Gann (1975) used rhodamine WT for karst tracing in
a limestone and dolomite system in Missouri. Three
fluorescent dyes (rhodamine B, rhodamine WT, and
fluorescein) were used by Brown and Ford (1971) in a
karst test in the Maligne Basin in Canada. The
highest recovery of dye (98%) was obtained for
141
-------
rhodamine WT. The fluorescein was not recovered at
all. Aulenbach et a/. (1978) compared rhodamine B,
rhodamine WT, and tritium as tracers in a delta sand.
The project involved tracing effluent from a sewage
treatment plant. The rhodamine B was highly
adsorbed, while the rhodamine WT and tritium yielded
similar breakthrough curves. Rhodamine WT seems
to be adsorbed less than rhodamine B or sulfo
rhodamine B (Table 7-3). Wilson (1971) found that
in column and field studies, rhodamine WT did show
sorptive tendencies.
Sulfo rhodamine B, also known as pontacyl brilliant
pink, is more expensive than the other rhodamine
dyes, and its toxicity appears to be slightly higher
than that of rhodamine WT. It has not been used
extensively as a ground-water tracer.
Blue fluorescent dyes have been used in increasing
amounts in the past decade in textiles, paper, and
other materials to enhance their white appearance.
Water which has been contaminated by domestic
waste can be used as a "natural" tracer if it contains
detectable amounts of the brighteners. Glover (1972)
described the use of optical brighteners in karst
environments. Examples of the brighteners are amino
G acid and photine CU. These two are the least
sensitive of the dyes reviewed (Table 7-2), but the
blue dyes have much lower background levels in
uncontaminated water than do the green or orange
dyes. Photine CU is significantly affected by
temperature variations, and both dyes are affected by
pH below 6.0. Amino G acid is fairly resistant to
adsorption.
Toxicity studies on optical brighteners were reviewed
by Akamatsu and Matsuo (1973). They concluded
that the brighteners do not present any toxic hazard
to man, even at excessive dosage levels.
7.8.5 Some Common Nonionized and Poorly
Ionized Compounds
A number of chemical compounds will dissolve in
water but will not ionize or will ionize only slightly
under normal conditions of pH and Eh found in
ground waters. Some of these compounds are
relatively difficult to detect in small concentrations,
others present a health hazard, and still others are
present in moderate to large concentrations in natural
waters, thus making the background effects difficult to
deal with in most settings. A list of a few of these
compounds is given in Table 7-4.
The use of these and similar compounds as injected
tracers in ground water is limited to rather special
cases. Of those listed, boric acid would probably act
most conservatively over long distances of ground-
water flow. Boric acid has been used successfully as
a tracer in a geothermal system (Downs ef a/., 1983).
Large concentrations, 1,000 mg/l or more, would
need to be used for injected tracers which,
unfortunately, would pose difficult environmental
questions if tracing were attempted in aquifers with
potable water. From the standpoint of health
concerns, sugars would be the most acceptable;
however, they decompose rapidly in the subsurface
and also tend to be sorbed on some materials.
Alcohols such as ethanol would also tend to be
sorbed on any solid organic matter which might be
present. Another problem with the use of most of
these compounds as tracers is that they would need
to be introduced in moderately large concentrations
which in turn would change the density and viscosity
(particularly for glycerin) of the injected tracer mixture.
Nevertheless, some of these compounds such as
sugars may be useful for simulating the movement of
other compounds which are also subject to rapid
decomposition but which are too hazardous to inject
directly into aquifers.
7.8.6 Gases
Numerous natural as well as artificially produced
gases have been found in ground water. Some of the
naturally produced gases can serve as tracers. Gas
can also be injected into ground water where it
dissolves and can serve as a tracer, but only a few
examples of it being used for ground-water tracers
are found in the literature. Gases of potential use in
hydrogeologic studies are listed in Table 7-5.
Inert Radioactive Gases. Chemically inert but
radioactive 133xe and 85Kr appear to be suitable for
many injected tracer applications (Robertson, 1969;
and Wagner, 1977), provided legal restrictions can be
overcome. Of the natural inert radioactive gases,
222Rn is the most abundant. It is one of the daughter
products from the spontaneous fission of 238u.
Radon is present in the subsurface, but owing to the
short half-life (3.8 d) of 222Rn, and the absence of
parent uranium nuclides in the atmosphere, radon is
virtually absent in surface water which has reached
equilibrium with the atmosphere. Surveys of radon in
surface streams and lakes have, therefore, been
useful in detecting the locations of places where
ground water enters surface waters (Rogers, 1958).
Inert Natural Gases. Because of their nonreactive and
nontoxic nature, noble gases are potentially useful
tracers. Helium is used widely as a tracer in industrial
processes. It also has been used to a limited extent
as a ground-water tracer (Carter et al, 1959). Neon,
krypton, and xenon are other possible candidates for
injected tracers because their natural concentrations
are very low (Table 7-5). Although the gases do not
undergo chemical reactions and do not participle in
ion exchange, the heavier noble gases (krypton and
xenon) do sorb to some extent on clay and organic
material.
The very low natural concentrations of noble gases in
ground water make them useful as tracers,
particularly in determining ground-water velocities in
regional aquifers. The solubility of the noble gases
142
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Table 7-4 Some Simple Compounds Which are Soluble in Water
Name Formula
Remarks
Silicic Acid
Boric Acid
Phosphoric Acid
Acetic Acid
Ethyl Alcohol
(Ethanol)
H,S,04
(After combination
with water)
H3B03
H3P04
C2H402
C2H60
Present in normal ground water in nonionized form in
concentrations of between 4 and 100 mg/L. Low toxicity.
Present in normal ground water in nonionized form in
concentrations of 0.05 to 2.0 mg/L. Toxic to plants above
1 to 5 mg/L. Toxic to humans in higher concentrations.
Ionizes above pH of 6.0. Will form complexes with other
dissolved constituents. Sorbs on or reacts with most
aquifer materials. Natural concentrations mostly between
0.05 mg/L and 0.5 mg/L.
Moderately toxic in high concentrations. Water soluble.
Natural concentrations are less than 0.1 mg/L in ground
water.
Major component of alcoholic drinks. Water soluble.
Natural concentrations are less than 0.05 mg/L in ground
water.
Sugars
Sucrose
Maltose
Lactose
Glucose
Glycerol
(Glycerin)
C12H220,i
C,2H22On
C6H1206
C3H604
Major components of human and animal foods. Water
soluble. Probably less than 0.2 mg/L in most ground
water.
Water soluble. Low toxicity. Probably absent in natural
ground water.
Table 7-5 Gases of Potential Use as Tracers
Approximate Natural Maximum Amount
Background Assuming in Solution Assuming
Equilibrium with 100% Gas at Pres-
Atmosphere at 20 °C sure of 1 atm at 20 °C
(mg gas/L water) (mg gas/L water)
Argon
Neon
Helium
Krypton
Xenon
Carbon Monoxide
Nitrous Oxide
0.57
1.7x10-
8.2 x 10-"
2.7 x 10-
5.7 x 10-5
6.0 x 10-'
3.3x10-
60.6
9.5
1.5
234
658
28
1100
decreases with an increase in temperature. The
natural concentrations of these gases in ground water
are, therefore, an indication of surface temperatures
at the time of infiltration of the water. This fact has
been used to reconstruct the past movement of water
in several aquifers (Sugisaki, 1969; Mazor, 1972;
Andrews and Lee, 1979).
Fluorocarbons. Numerous artificial gases have been
manufactured during the past decade and several of
these gases have been released in sufficient volumes
to produce measurable concentrations in the
atmosphere on a worldwide scale. One of the most
interesting groups of these gases are the
fluorocarbons (Table 7-6). The gases generally pose
a very low biological hazard, they are generally stable
for periods measured in years, they do not react
chemically with other materials, they can be detected
in very low concentrations, and they sorb only slightly
on most minerals. They do sorb strongly, however, on
organic matter.
Fluorocarbons have two primary applications. First,
because large amounts of fluorocarbons were not
released into the atmosphere until the later 1940s and
early 1950s, the presence of fluorocarbons in ground
water indicates that the water was in contact with the
atmosphere within the past 30 to 40 yr and that the
ground water is very young (Thompson and Hayes,
1978). The second application of fluorocarbon
compounds is for injected tracers (Thompson, Hayes,
and Davis, 1974). Because detection limits are so
low, large volumes of water can be labeled with the
tracers at a rather modest cost. Despite the problem
of sorption on natural material and especially on
organics, initial tests have been quite encouraging.
Stable Isotopes. An isotope is a variation of an
element produced by differences in the number of
neutrons in the nucleus of that element. However, the
difficulty in detecting small artificial variations of most
isotopes against the natural background, the high cost
of their analysis, and the expense of preparing
isotopically enriched tracers, means that stable
143
-------
Table 7-6 Properties of Fluorocarbon Compounds
Common Name
Freon-1 1
Freon-12
Freon-1 13
Chemical
Formula
CCI3F
CCI2F2
CCI2F-CCIF2
CBrCIF2
CBr2F2
CBrl-CBrF2
Boiling Point
at 1 atm ( °C)
23.8
-29.8
47.6
-4.0
24.5
47.3
Solubility in Water
at 25 °C (weight %)
0.11
0.028
0.017
unknown
unknown
unknown
isotopes are rarely used for artifically injected tracer
studies in the field.
Research into the topic of stable isotopes of various
elements in natural waters is progressing rapidly, and
the potential usefulness of these isotopes to ground-
water tracing will undoubtedly increase markedly in
the near future.
The most common use of studies of 2H and 180 has
been to trace the large-scale movement of ground
water and to locate areas of recharge (Gat, 1971;
Fritz and Fontes, 1980; and Ferronsky and Polyakov,
1982).
The two abundant isotopes of nitrogen (14N and 15N)
can vary significantly in nature. Ammonia escaping as
vapor from decomposing animal wastes, for example,
will tend to remove the lighter (14N) nitrogen and will
leave behind a residue rich in heavy nitrogen. In
contrast, many fertilizers with an ammonia base will
be isotopically light. Natural soil nitrate will be
somewhat between these two extremes. As a
consequence, nitrogen isotopes have been useful in
helping to determine the origin of unusually high
amounts of nitrate in ground water. Also, the
presence of more than about 5 mg/l of nitrate
commonly is an indirect indication of contamination
from chemical fertilizers and sewage.
The stable sulfur isotopes (32S, 34s, and 36$) have
been used to distinguish sulfate originating from
natural dissolution of gypsum (CaS04*2H20) from
sulfate originating from an industrial spill of sulfuric
acid (H2SO4).
Two stable isotopes of carbon (12C and 13Q and one
unstable isotope 04C) are used in hydrogeologic
studies. Most isotopic studies of carbon in water have
been centered on 14C which will be discussed in a
later portion of this chapter. Although not as
commonly studied as 14C, the ratio of the stable
isotopes, 13C/12C, is potentially useful in sorting out
the origins of certain contaminants found in water. For
example, methane (CH4) originating from some deep
geologic deposits is isotopically heavier then methane
originating from near surface sources. This contrast
forms the basis for identifying aquifers contaminated
with methane from pipelines and from subsurface
storage tanks.
Isotopes of other elements such as chlorine,
strontium, and boron are related more to the
determination of regional directions of ground-water
flow than to problems of the identification of sources
of contamination.
Radionuclides. Radioactive isotopes of various
elements are collectively referred to as radionucl ides.
In the early 1950's there was great enthusiasm for
using radionuclides both as natural "environmental"
tracers and as injected artificial tracers. The use of
artificially injected radionuclides has all but ceased in
many countries, including the United States. Most use
of artificially introduced radioactive tracers is confined
to carefully controlled laboratory experiments or to
deep petroleum production zones which are devoid of
potable water. However, the environmental use has
been expanded greatly until it is now a major
component of many hydrochemical studies.
A number of radionuclides are present in the
atmosphere from natural and artificial sources. Many
of these are carried into the subsurface by rain water.
The most common hydrogeologic use of these
radionuclides is to obtain some estimate of the
average length of time ground water has been
isolated from the atmosphere. This is complicated by
dispersion in the aquifer and mixing in wells that
sample several hydrologic zones. Nevertheless, it can
usually be established that most or virtually all of the
ground water is older than some given limiting value.
In many situations we can say, based on atmospheric
radionuclides, that the ground water was recharged
more than 1,000 years ago or that, in another region,
all the ground water in a given shallow aquifer is
younger than 30 years.
Since the 1950s, atmospheric tritium, the radioactive
isotope of hydrogen (3H), has been dominated by
144
-------
tritium from the detonation of thermonuclear devices.
Thermonuclear explosions had increased the
concentration of tritium in local rainfall to more than
1,000 TU in the northern hemisphere by the early
1960s (Figure 7-14). As a result, ground water in the
northern hemisphere which has more than about 5
TU is generally less than 30 years old. Very small
amounts, 0.05 to 0.5 TU, can be produced by natural
subsurface processes, so the presence of these low
levels does not necessarily indicate water 40 to 60
years old or small amounts of more recent water
mixed with very old water.
Figure 7-14 Average annual tritium concentration of rainfall
and snow for Arizona, Colorado, New Mexico,
and Utah.
4000
3500
3000
2500
2000
1500
1000
500
0
Table 7-7 Commonly Used Radioactive Tracers for
Ground-Water Studies
56 58 60 62 64
70 72 74 76 78 80 82
Year
The radioactive isotope of carbon, 14C, is also widely
studied in ground water. In practice, the use of 14C is
rarely simple. Sources of old carbon, primarily from
limestone and dolomite, will dilute the sample. A
number of processes, such as the formation of CH4
gas or the precipitation of carbonate minerals, will
fractionate the isotopes and alter the apparent age.
The complexity of the interpretation of 14C "ages" of
water is so great that it should be attempted only by
hydrochemists specializing in isotope hydrology.
Despite the complicated nature of 14C studies, they
are highly useful in determining the approximate
residence time of old water (500 to 30,000 yr) in
aquifers. In certain circumstances, this information
cannot be obtained in any other way.
A number of radionuclides commonly used as tracers
are shown in Table 7-7
Radionuclide
JH
MP
51 Cr
MCo
B2Br
85Kr
131 1
'MAu
Half-Life
y = year,
d = day,
h = hour
12.3y
14.3d
27.8d
5.25y
33.4h
10.7y
8.1d
2.7d
Chemical Compound
H20
Na2HP04
EDTA-Cr and CrCI3
EDTA-Co and K3Co (CN.)
NH4Br, NaBr, LiBr
Kr (gas)
I and Kl
AuCI3
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Sorey, M.L. 1971. Measurement of Vertical Ground-
Water Velocity from Temperature Profiles in Wells.
Water Resources Research 7(4):963-970.
Sugisaki, R. 1969. Measurement of Effective Flow
Velocity of Groundwater by Means of Dissolved
Gases. American Journal Science 259:144-153.
Tennyson, L.C., and C. D. Settergren. 1980.
Percolate Water and Bromide Movement in the Root
Zone of Effluent Irrigation Sites. Water Resources
Bulletin 16(3):433-437.
Tester, J.W., R. L. Bivens, and R. M. Potter. 1982.
Interwell Tracer Analysis of a Hydraulically Fractured
Grantitic Geothermal Reservoir. Society of Petroleum
Engineers Journal 8:537-554.
Theis, C.V. 1963. Hydrologic Phenomena Affecting
the Use of Tracers in Timing Ground-Water Flow:
Radioisotopes in Hydrology. International Atomic
Energy Agency (Tokyo Symposium), Vienna, Austria.
Thompson, G.M., and J. M. Hayes. 1978.
Trichlorofluoromethane in Ground Water. A Possible
Tracer and Indicator of Ground-Water Age. Water
Resources Research 15(3):546-554.
Thompson, G.M., J. M. Hayes, and S. N. Davis.
1974. Fluorocarbon Tracers in Hydrology.
Geophysical Research Letters 1:177-180.
Todd, O.K. 1980. Groundwater Hydrology , 2nd ed.
John Wiley and Sons, New York, NY.
Wagner, O.R. 1977. The Use of Tracers in
Diagnosing Interwell Reservoir Heterogeneities. Jour.
Petroleum Technology 11:1410-1416.
Wilson, J.F. 1968. Fluorometric Procedures for Dye
Tracing. In: Techniques of Water-Resources
Investigations of the U.S. Geological Survey.
Wilson, L.G. 1971. Investigations on the Subsurface
Disposal of Waste Effluent at Inland Sites. Final report
to Office of Saline Water, Grant U14-01-0001-
1805. Water Resources Research Center, Tucson,
AZ.
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Ground Water 16(6):398-403.
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CHAPTER 8
THE USE OF MODELS IN MANAGING GROUND-WATER PROTECTION PROGRAMS
8.1 The Utility of Models
8.1.1 Introduction
Mathematical models rely on the quantification of
relationships between specific parameters and
variables to simulate the effects of natural processes
(Figures 8-1, 8-2). As such, mathematical models
are abstract and provide little in the way of a directly
observable link to reality. Despite this lack of intuitive
grace, mathematical models can generate powerful
insights into the functional dependencies between
causes and effects in the real world. Large amounts
of data can be generated quickly, and experimental
modifications can be made with minimal effort, so that
many possible situations can be studied in great detail
for a given problem.
Figure 8-1 Typical ground-water contamination scenario.
Several water-supply production wells are
located downgradient of a contaminant source.
The geology is complex.
Figure 8-2 Possible contaminant transport model grid
design for the situation shown in Figure 8-1.
Values for natural process parameters would be
specified at each node of the grid in performing
simulations. The grid density is greatest at the source
and at potential impact locations.
8.1.2 Management Applications
Mathematical models can and have been used to help
organize the essential details of complex ground-
water management problems so that reliable solutions
are obtained (Holcomb Research Institute, 1976;
Bachmat et a/., 1978; U.S. Congress, 1982; van der
Heijde et a/., 1985). Some of the principal areas
where mathematical models are now being used to
assist in the management of ground-water protection
programs are:
o Appraising the physical extent, and chemical and
biological quality, of ground-water reservoirs
(e.g., for planning purposes)
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o Assessing the potential impact of domestic,
agricultural, and industrial practices (e.g., for
permit issuance)
o Evaluating the probable outcome of remedial
actions at waste sites, and aquifer restoration
techniques generally
o Providing health-effects exposure estimates.
The success of these efforts depends on the
accuracy and efficiency with which the natural
processes controlling the behavior of ground water,
and the chemical and biological species it transports,
are simulated (Boonstra and de Ridder, 1976; Mercer
and Faust, 1981; Wang and Anderson, 1982). The
accuracy and efficiency of the simulations, in turn, are
heavily dependent on subjective judgments made by
the modeler and management.
In the current philosophy of ground-water protection
programs, the value of a ground-water resource is
bounded by the most beneficial present and future
uses to which it can be put (U.S. EPA, 1984). In most
instances, physical appraisals of ground-water
resources are conducted within a framework of
technical and economic classification schemes.
Classification of entire ground-water basins by
potential yield is a typical first step (Domenico, 1972).
After the initial identification and evaluation of a
ground-water resource, strategies for its rational
development need to be devised.
Development considerations include the need to
protect vulnerable recharge areas, and the possibility
of conjunctive use with available surface waters
(Kazmann, 1972). Ground-water rights must be fairly
administered to assure adequate supplies for
domestic, agricultural, and industrial purposes.
Because basinwide or regional resource evaluations
normally do not provide sufficient resolution for water
allocation purposes, more detailed characterizations
of the properties and behavior of an aquifer, or of a
subdivision of an aquifer, are usually needed. Hence,
subsequent classifications may involve local
estimation of net annual recharge, rates of outflow,
and the pumpage which can be sustained without
undesirable effects.
The consequences of developments which might
affect ground-water quality may be estimated initially
by employing generalized classification schemes; for
example, classifications based on regional
hydrogeologic settings have been presented (Health,
1982; Aller et a/., 1985). Very detailed databases,
however, must be created and molded into useful
formats before decisions can be made on how best to
protect and rehabilitate ground-water resources from
site-specific incidents of natural and manmade
contamination.
The latter are ordinary ground-water management
functions which benefit from the use of mathematical
models. There are other uses, however, which ought
to be considered by management. The director of the
International Ground Water Modeling Center
discussed the role of modeling in the development of
ground-water protection policies recently, noting its
success in many policy formulation efforts in the
Netherlands, the United States, and Israel.
Nevertheless, he concluded that modeling was not
widely relied upon for decision-making by managers.
The primary obstacle has been an inability by
modelers and program managers to communicate
effectively (van der Heijde, 1985). The top executives
of a leading high-tech ground-water contamination
consulting firm made the same point clearly, going on
to highlight the need for qualified personnel
appreciative of the appropriateness, underlying
assumptions, and limitations of specific models (Faust
et a/., 1981). Because these views are widely held by
technical professionals, it will be emphasized herein
that mathematical models are useful only within the
context of the assumptions and simplifications on
which they are based. If managers are mindful of
these factors, however, mathematical models can be
a tremendous asset in the decision-making process.
8.7.3 Modeling Contamination Transport
Associated with most hazardous waste sites is a
complex array of chemical wates and the potential for
ground-water contamination. Since the
hydrogeologic settings of these sites are usually quite
complicated data acquisition and interpretation
methods are needed which can examine to an
unprecedented degree the physical, chemical, and
biological processes which control the transport and
fate of ground-water contaminants. The methods
and tools that have been in use for large-scale
characterizations (e.g., regional water quality studies)
are applicable in concept to the specialized needs of
hazardous waste site investigations; however, the
transition to local-scale studies is not without
scientific and economic consequences. In part, this
stems from the highly variable nature of contaminant
distributions at hazardous waste sites; but it also
results from the limitations of the methods, tools, and
theories used. Proper acknowledgement of the
inherent limitations means that one must project the
consequences of their use within the framework of
the study at hand.
Assessments of the potential for contaminant
transport require interdisciplinary analyses and
interpretations. Integration of geologic, hydrologic,
chemical, and biological approaches into an effective
contaminant transport evaluation can only be possible
if the data and concepts invoked are sound. The data
must be accurate, precise, and appropriate for the
intended problem scale. Just because a given
parameter (e.g., hydraulic conductivity) has been
measured correctly at certain points with great
reproducibility, is no guarantee that those estimates
represent the volumes of aquifer material assigned to
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them by a modeler. The degree to which the data are
representative, therefore, is not only relative to the
physical scale of the problem, it is relative to the
conceptual model to be used for interpretation efforts.
It is crucial, then, to carefully define and qualify the
conceptual model. In so doing, special attention
should be given to the possible spatial and temporal
variations of the data that will be collected.
To circumvent the impossibly large numbers of
measurements and samples which would be needed
to eliminate all uncertainties regarding the true
relationships of parameters and variables, more
comprehensive theories are constantly under
development. The use of newly developed theories to
help solve field problems, however, is often a
frustrating exercise. Most theoretical advances call for
some data which are not yet practically obtainable
(e.g., chemical interaction coefficients, relative
peremeabilities of immiscible solvents and water, and
so on). The "state-of-the-art" in contaminant
transport assessments is necessarily a compromise
between the sophistication of "state-of-the-
science" theories, the current limitations regarding
the acquisition of specific data, and economics. In
addition, the best attempts to obtain credible data still
fall prey to natural and anthropogenic variabilities; and
these lead to the need for considerable judgment on
the part of the professional.
Despite these limitations, how well the problem is
conceptualized remains the most serious concern in
modeling efforts. For example, researchers recently
produced dramatic evidence to show that
extrapolations of two-dimensional model results to a
truly three-dimensional problem lead to wildly
inaccurate projections of the actual behavior of the
system under study (Molz et a/., 1983). Therefore it is
incumbent on model users to recognize the difference
between an approximation and a misapplication.
Models should never be used strictly on the basis of
familiarity or convenience; an appropriate model
should always be sought.
8.1.4 Categories of Models
The foregoing is not meant to imply that appropriate
models exist for all ground-water problems, because
a number of natural processes have yet to be fully
understood. This is especially true for ground-water
contaminant transport evaluations, where the
chemical and biological processes are still poorly
defined. For, although great advances have been
made concerning the behavior of individual
contaminants, studies of the interactions between
contaminants are still in their infancy. Even the
current understanding of physical processes lags
behind what is needed, such as in the mechanics of
multiphase flow and flow through fractured rock
aquifers. Moreover, certain well-understood
phenomena pose unresolved difficulties for
mathematical formulations, such as the effects of
partially penetrating wells in unconfined aquifers.
The technical-use categories of models are varied,
but they can be grouped as follows (Bachmat et a/.,
1978; van der Heijde et a/., 1985):
o Parameter identification models
o Prediction models
o Resource management models
o Data manipulation codes.
Parameter identification models are most often used
to estimate the aquifer coefficients determining fluid
flow and contaminant transport characteristics, like
annual recharge (Puri, 1984), coefficients of
permeability and storage (Shelton, 1982; Khan, 1986a
and 1986b), and dispersivity (Guven ef a/., 1984;
Strecker and Chu, 1986). Prediction models are the
most numerous kind of model, and abound because
they are the primary tools for testing hypotheses
about the problem one wishes to solve (Andersen et
a/., 1984; Mercer and Faust, 1981; Krabbenhoft and
Anderson, 1986).
Resource management models are combinations of
predictive models, constraining functions (e.g., total
pumpage allowed) and optimization routines for
objective functions (e.g., optimization of wellfield
operations for minimum cost or minimum
drawdown/pumping lift). Very few of these are so well
developed and fully supported that they may be
considered practically useful, and there does not
appear to be a significant drive to improve the
situation (van der Heijde, 1984a and 1984b; van der
Heijde et a/., 1985). Data manipulation codes also
have received little attention until recently. They are
now becoming increasingly popular, because they
simplify data entry ("preprocessors") to other kinds of
models and facilitate the production of graphic
displays ("postprocessors") of the data outputs of
other models (van der Heijde and Srinivasan, 1983;
Srinivasan, 1984; Moses and Herman, 1986). Other
software packages are available for routine and
advanced statistics, specialized graphics, and
database management needs (Brown, 1986).
8.2 Assumptions, Limitations, and Quality
Control
The many natural processes that affect chemical
transport from point to point in the subsurface can be
arbitrarily divided into three categories: physical,
chemical, and biological (Table 8-1). Conceptually,
contaminant transport in the subsurface is an
undivided phenomenon composed of these processes
and their interactions. At this level the transport
process may be gestalt: the sum of its parts,
measured separately, may not equal the whole
because of interactions between the parts. In the
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theoretical context, a collection of scientific laws and
empirically derived relationships comprise the overall
transport process. The universally shared premise
that underlies theoretical expressions is that there are
no interactions, measurable or otherwise.
Table 8-1 Natural processes that affect subsurface
contaminant transport.
Physical Processes
Advection (porous media velocity)
Hydrodynamic dispersion
Molecular diffusion
Density stratification
Immiscible phase flow
Fractured media flow
Chemical Processes
Oxidation —reduction reactions
Radionuclide decay
Ion—exchange
Complexation
Co—solvation
Immiscible phase partitioning
Sorption
Biological Processes
Microbial population dynamics
Substrate utilization
Biotransformation
Adaption
Co —metabolism
Significant errors may result from the discrepancy
between conceptual and theoretical approaches. Also
the simplifications of theoretical expressions used to
solve practical problems can cause substantial errors
in the most careful analyses. Assumptions and
simplifications, however, must often be made in order
to obtain mathematically tractable solutions. Because
of this, the magnitude of errors that arise from each
assumption and simplification must be carefully
evaluated. The phrase magnitude of errors is
emphasized because highly accurate evaluations
usually are not possible. Even rough approximations
are rarely trivial exercises because they frequently
demand estimates of some things which are as yet
ill-defined.
8.2.7 Physical Processes
Until recently, ground-water scientists studied
physical processes to a greater degree than chemical
or biological processes. This bias resulted in large
measure from the fact that, in the past, ground-
water practitioners dealt mostly with questions of
adequate water supplies. As quality considerations
began to dominate ground-water issues, the need
for studies of the chemical and biological factors, as
well as more detailed representations of the physical
factors became apparent.
There are two complimentary ways to view the
physical processes involved in subsurface
contaminant transport: the piezometric (pressure)
viewpoint and the hydrodynamic viewpoint. Ground-
water problems of yesterday could be addressed by
the former, such as solving for the change in
pressure head caused by pumping wells.
Contamination problems of today also require detailed
analyses of wellfield operations, for example, pump-
and-treat plume removals; however, solutions
depend principally on hydrodynamic evaluations, such
as computing ground-water velocity (advection)
distributions and dispersion estimates for migrating
plumes.
8.2.1.1 Advection and dispersion
Ground-water velocity distributions can be
approximated if the variations in hydraulic
conductivity, porosity, and the strength and location of
recharge and discharge can be estimated. While
there are several field and laboratory methods for
estimating hydraulic conductivity, these are not
directly comparable because different volumes of
aquifer material are affected by different tests.
Laboratory permeameter tests, for example, obtain
measurements from small core samples and thus
give point value estimates. These tests are generally
reliable for consolidated rock samples, such as
sandstone, but can be highly unreliable for
unconsolidated samples, such as sands, gravels, and
clays. Pumping tests give estimates of hydraulic
conductivity that are averages over the entire volume
of aquifer subject to the pressure changes induced by
pumping. These give repeatable results, but they are
often difficult to interpret. Tracer tests are also used
to estimate hydraulic conductivity in the field, but are
difficult to conduct properly.
Regardless of the estimation technique used, the best
that can be expected is order-of-magnitude
estimates for hydraulic conductivity at the field scale
appropriate for site-specific work. Conversely,
porosity estimates that are accurate to better than a
factor of two can be obtained. Estimation of the
strength of nonpoint sources of recharge to an
aquifer, such as infiltrating rainfall and leakage from
other aquifers, is another order-of-magnitude effort.
Similarly, nonpoint sources of discharge, such as
losses to gaining streams, are difficult to quantify.
Estimation of the strength of point sources of
recharge or discharge (injection or pumping wells)
can be highly accurate.
Consequently, it is not possible to generalize the
quality of velocity distributions. They may be accurate
to within a factor of two for very simple aquifers, but
are more often accurate to an order-of-magnitude
only. This situation has changed little over the past 20
years because better field and laboratory methods for
characterizing velocity distributions have not been
developed. This, however, is not the primary difficulty
associated with defining the advective part of
contaminant transport in the subsurface. The primary
difficulty is that field tests for characterizing the
physical parameters that control velocity distributions
are not incorporated into contamination investigations
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on a routine basis. The causes seem to be: a
perception that mathematical models can "back-out"
an approximation of the velocity distribution
(presumably eliminating the need for field tests);
unfamiliarity with such methods by many practitioners;
and a perception that field tests are too expensive. A
more field-oriented approach is preferable because
the non-uniqueness of modeling results has been
amply demonstrated, and this leads to uncertain
decisions regarding the design of remedies.
Dispersion estimates are predicted on velocity
distribution estimates and their accuracy is therefore
directly dependent on the accuracy of the estimated
hydraulic conductivity distribution. Tracer tests have
been the primary method used to determine
dispersion coefficients until recently. Presently there
are suggestions that any field method capable of
generating a detailed understanding of the spatial
variability of hydraulic conductivity, which in turn
could give an accurate representation of the velocity
distribution, may be used to derive estimates of
dispersion coefficients. The manner in which data
from field tests should be used to derive estimates of
dispersion coefficients, however, is a controversial
issue. There are both deterministic and stochastic
schools of thought, and neither has been conclusively
demonstrated in complex hydrogeological settings.
8.2. t.2 Complicating factors
Cert ain subtleties of the spatial variability of hydraulic
conductivity must be understood because of its key
role in the determination of velocity distributions and
dispersion coefficients. Hydraulic conductivity is also
known as the coefficient of permeability because it is
comprised of fluid factors as well as the intrinsic
permeability of the stratum in question. This means
that a stratum of uniform intrinsic permeability (which
depends strictly on the arrangement of its pores) may
have a wide range of hydraulic conductivity because
of differences in the density and viscosity of fluids
that are present. The result is a dramatic downward
shift in local flow directions near plumes that have as
little as a one percent increase in density relative to
uncontaminated water. Such density contrasts
frequently occur at landfills and waste impoundments.
It is often necessary to correct misimpressions of the
direction of a plume because density considerations
were not addressed.
Many solvents and oils are highly insoluble in water,
and may be released to the subsurface in amounts
sufficient to form a separate fluid phase. Because that
fluid phase will probably have viscosity and density
different from freshwater, it will flow at a rate and,
possibly, in a direction different from that of the
freshwater with which it is in contact. If an immiscible
phase has a density approximately the same or less
than that of ground water, this phase will not move
down past the capillary fringe of the ground water.
Instead, it will flow along the top of the capillary fringe
in the direction of the maximum water-level elevation
drop. If the density of an immiscible phase is
substantially greater than the ground water, the
immiscible phase will push its way into the ground
water as a relatively coherent blob. The primary
direction of its flow will then be down the dip of the
first impermeable stratum encountered. There is a
great need for better means of characterizing such
behavior for site-specific applications. Currently,
estimation methods are patterned after multiphase oil
reservoir simulators. One of the key extensions
needed is the ability to predict the transfer of trace
levels of contaminants, such as xylenes from
gasoline, from the immiscible fluid to ground water.
Anisotropy is a subtlety of hydraulic conductivity
which relates to structural trends of the rock or
sediments of which an aquifer is composed.
Permeability and hydraulic conductivity are
directionally dependent in anisotropic strata. When
molten material from deep underground crystallizes to
form granitic or basaltic rocks, for instance, it forms
cleavage planes which may later becomes the
preferred directions of permeability. Marine sediments
accumulate to form sandstone, limestone, and shale
sequences that have much less vertical than
horizontal permeability. The seasonal differences in
sediments that accumulate on lakebeds, and the
stratification of grain sizes deposited by streams as
they mature, give rise to similar vertical-to-
horizontal anisotropy. Streams also cause anisotropy
within the horizontal plane, by forming and reworking
their sediments along a principal axis of movement.
These structural variations in permeability would be of
minimal concern except that ground water does not
flow at right angles to water-level elevation contours
under anisotropic conditions. Instead, flow proceeds
along oblique angles, with the degree of deviation
from a right-angle pathway proportional to the
amount of anisotropy. This fact is all too often ignored
and the causes again seem to be a reluctance to
conduct the proper field tests, combined with an
over-reliance on mathematical modeling.
If the pathways created by cleavage planes and
fractures begin to dominate fluid flow through a
subsurface stratum, the directions and rates of flow
are no longer predictable by the equations used for
porous rock and sediments. There have been a
number of attempts to represent fractured flow as an
equivalent porous medium, but these tend to give
poor predictions when major fractures are present
and when there are too few fractures to guarantee a
minimum degree of interconnectedness. Other
representations that have been studied are various
dual porosity models, in which the bulk matrix of the
rock has one porosity and the fracture system has
another. Further development of the dual porosity
approach is limited by the difficulty in determining a
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transfer function to relate the two different porosity
schemes.
8.2.1.3 Considerations for predictive modeling
Equations for the combined advection-dispersion
process are used to estimate the time during which a
nonreactive contaminant will travel a specific
distance, the pathway it will travel, and its
concentration at any point. The accuracy of most
predictions is only fair for typical applications,
because of the complexity of the problems and the
scarcity of site-specific hydrogeologic data. The lack
of such data can be improved with much less effort
than is commonly presumed, especially when costs of
another round of chemical sampling are compared
with the costs of additional borings, core retrievals,
geophysical logging, or permeability testing.
Equations that assume a nonreactive contaminant
have limited usefulness, because most contaminants
react with other chemical constituents in subsurface
waters and with subsurface solids in a manner that
affects the rate at which they travel. Nevertheless,
nonreactive advection-dispersion equations are often
used to generate "worst-case" scenarios, on the
presumption that the maximum transport velocity is
obtained (equal to that of pure water). This may not
be as useful as it first seems. Remedial action
designs require detailed breakdowns of which
contaminants will arrive at extraction wells and when;
how long contaminants will continue their slow
release from subsurface solids; and whether the
contaminants will be transformed into other chemical
species by chemical or biological forces. To address
these points, special terms must be added to the
advection-dispersion equations.
8.2.2 Chemical Processes
As difficult as the foregoing complications may be,
predicting how chemical contaminants move through
the subsurface is a relatively trivial matter when the
contaminants behave as ideal, nonreactive
substances. Unfortunately, such behavior is limited to
a small group of chemicals. The actual situation is
that most contaminants will, in a variety of ways,
interact with their environment through biological or
chemical processes.
This section focuses on the dominant chemical
processes that may ultimately affect the transport
behavior of a contaminant. As with the physical
processes previously discussed, some of the
knowledge of chemical processes has been
translated into practical use in predictive models.
However, the science has, in many instances,
advanced well beyond what is commonly practiced.
Furthermore, there is considerable evidence that
suggests that numerous undefined processes affect
chemical mobility. Most of the deviation from ideal
nonreactive behavior of contaminants relates to their
ability to change physical form by energetic
interactions with other matter. The physical-chemical
interactions may be grouped into: alterations in the
chemical or electronic configuration of an element or
molecule; alterations in nuclear composition; the
establishment of new associations with other
chemical species; and, interactions with solid
surfaces.
8.2.2.1 Chemical/electronic alterations
The first of these possible changes is typified by
oxidation-reduction or redox reactions. This class of
reactions is especially important for inorganic
compounds and metallic elements because the
reactions often result in changes in solubility,
complexing capacity, or sorptive behavior, which
directly impact the mobility of the chemical. Redox
reactions are reasonably well understood, but there
are practical obstacles to applying the known science
because it is difficult to determine the redox state of
the aquifer zone of interest and to identify and
quantify the redox-active reactants.
Hydrolysis, elimination, and substitution reactions that
affect certain contaminants also fit into this
classification. The chemistry of many organic
contaminants has been well defined in surface water
environments. The influence of unique aspects of the
subsurface, not the least of which is long residence
time, on such transformations of important organic
pollutants is currently under investigation. There is
also a need to investigate the feasibility of promoting
in-situ abiotic transformations that may enhance the
potential for biological mineralization of pollutants.
8.2.2.2 Nuclear alterations
Another chemical process interaction, which results in
internal rearrangement of the nuclear structure of an
element, is well understood. Radiodecay occurs by a
variety of routes, but the rate at which it occurs is
always directly proportional to the number of
radioactive atoms present. This fact seems to make
mathematical representation in contaminant transport
models quite straightforward because it allows
characterization of the process with a unique, well
defined decay constant for each radionuclide.
A mistake that is often made when the decay
constant is used in models involves the physical form
of the reactant. If the decay constant is applied to the
fluid concentrations and no other chemical
interactions are allowed, then incorporation of the
constant in the subroutine which computes fluid
concentrations will not cause errors. If the situation
being modeled involves chemical interactions such as
precipitation, ion-exchange, or sorption, which
temporarily remove the radionuclide from solution,
then it is important to use a second subroutine to
account for the non-solution phase decay of the
radionuclide.
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8.2.2.3 Chemical associations
The establishment of new associations with other
chemical species is not as well understood. This
category includes ion-exchange, complexation, and
co-solvation. The lack of understanding derives from
the nonspecific nature of these interactions, which
are, in many instances, not characterized by the
definite proportion of reactants to products
(stoichiometry) typical of redox reactions. While the
general principles and driving mechanisms by which
these interactions occur are known, the complex
subsurface matrix in which they occur provides many
possible outcomes and renders predictions uncertain.
Ion-exchange and complexation reactions heavily
influence the mobility of metals and other ionic
species in the subsurface in a reasonably predictable
fashion. Their influence on organic contaminant
transport, however, is not well understood. Based on
studies of pesticides and other complex organic
molecules, natural organic matter (such as humic and
fulvic materials) can complex and thereby enhance
the apparent solubility and mobility of synthetic
organic chemicals. Research is needed to define the
magnitude of such interactions, not only with naturally
occurring organic molecules but also with man-made
organics present in contaminated environments.
Research is also needed to determine if these
complexes are stable and liable to transport through
the subsurface. Examination of the degree to which
synthetic organic chemicals complex toxic metals is
also necessary. There is no theoretical objection to
such interactions, and there is ample evidence that
metals are moving through the subsurface at many
waste sites.
Co-solvation occurs when another solvent is in the
aqueous phase at concentrations that enhance the
solubility of a given contaminant. This occurs in
agricultural uses, for example, where highly insoluble
pesticides and herbicides are mixed with organic
solvents to increase their solubility in water prior to
field application. There is every reason to expect
similar behavior at hazardous waste sites, where a
variety of solvents are typically available. At present,
prediction of the extent of the solubility increases that
might occur at disposal sites in the complex mixture
of water and organic solvents is essentially
impossible. Researchers have started examining co-
solvation as an influence on pollutant transport, by
working on relatively simple mixed solvent systems.
This research will be extremely useful, even if the
results are limited to a qualitative appreciation for the
magnitude of the effects.
At the extreme, organic solvents in the subsurface
may result in a phase separate from the aqueous
phase. In addition to movement of this separate
phase through the subsurface, contaminant mobility
that involves partitioning of organic contaminants
between the organic and aqueous phases must also
be considered. The contaminants will move with the
organic phase and will, depending on aqueous phase
concentrations, be released into the aqueous phase
to a degree roughly proportional to their octanol-
water partition coefficients. An entire range of effects
is possible, from increasing to slowing the mobility of
the chemical in the subsurface relative to its migration
rate in the absence of the organic phase. The
equilibrium partitioning process increases the total
volume of ground water affected by contaminants, by
releasing a portion of the organic phase constituents
into adjacent waters. It may also interfere with
transformation processes by affecting pollutant
availability for reaction, or by acting as a biocidal
agent to the native microflora.
8.2.2.4 Surface interactions
Of those interactions that involve organic chemicals in
the environment, none has been as exhaustively
studied as sorption. Sorption studies relate, in terms
of a sorption isotherm, the amount of contaminant in
solution to the amount associated with the solids.
Most often the sorption term in transport models is
estimated for simplicity from the assumption that the
response is linear. This approximation can produce
serious mass balance errors. Typically, the
contaminant mass in the solution phase is under-
estimated and contaminant retardation is thereby
over-estimated. In practical applications, this means
that the contaminant can be detected at a monitoring
well long before it is anticipated. To resolve the
discrepancy between predicted and actual transport,
most practitioners arbitrarily adjust some other
poorly-characterized model parameter, for example,
dispersion. This leads to the creation of a model that
does not present various natural process influences in
proper perspective. The predictions from such models
are likely to be qualitatively, as well as quantitatively,
incorrect. More widespread consideration should be
given to accurate representation of non-linear
sorption, particularly in transport modeling at
contaminated sites.
The time dependency of the sorption process is a
related phenomenon that has also been largely
ignored in practical applications of sorption theory.
Most models assume that sorption is instantaneous
and completely reversible. A growing body of
evidence argues to the contrary, not only for large
organic molecules in high-carbon soils and
sediments, but also for solvent molecules in low-
carbon aquifer materials. Additionally, there must be
some subtle interplay between sorption kinetics and
ground-water flow rates which gains significance in
pump-and-treat remediation efforts, where flow
rates are routinely substantially increased. Constant
pumpage at moderate-to-high flow rates may not
allow contaminants that are sorbed to solids sufficient
times of release to increase solution concentrations to
maximum (equilibrium) levels prior to their removal
from the aquifer. Hence, treatment costs may rise
155
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substantially due to the prolonged pumping required
to remove all of the contaminants and due to the
lowered efficiency of treatment of the less
contaminated pumped waters.
Evidence from Superfund sites and ongoing research
activities suggests that contaminant association with a
solid surface does not preclude mobility. In many
instances, especially in glacial tills that contain a wide
distribution of particle sizes, fine aquifer materials
have accumulated in the bottom of monitoring wells.
Iron-based colloids have been identified in ground
water downgradient from a site contaminanted with
domestic wastewater. If contaminants can associate
with these fine particles, their mobility through the
subsurface could be markedly enhanced. To
determine the significance of particle transport to
pollutant movement, studies must be performed at
such contaminanted sites.
Although knowledge about chemical processes that
function in the subsurface has been significantly
expanded in recent years, this information is only
slowly finding its way into practical interpretations of
pollutant transport at contaminated sites. Evidence
from field sites suggests that much remains to be
learned about these processes.
8.2.3 Biological Processes
Many contaminants that enter the subsurface
environment are biologically reactive. Under
appropriate circumstances they can be completely
degraded to harmless products. Under other
circumstances, however, they can be transformed to
new substances that are more mobile or more toxic
than the original contaminant. Quantitative predictions
of the fate of biologically reactive substances are at
present very primitive, particularly compared to other
processes that affect pollutant transport and fate. This
situation resulted from the ground-water
community's choice of an inappropriate
conceptualization of the active processes: subsurface
biotransformations were presumed to be similar to
biotransformations known to occur in surface water
bodies. Only very recently has detailed field work
revealed the inadequacy of the traditional view.
8.2.3.1 Surface water model analogy
As little as five years ago ground-water scientists
considered aquifers and soils below the zone of plant
roots to be essentially devoid of organisms capable of
transforming contaminants. As a result, there was no
reason to include terms for biotransformations in
transport models. Recent studies have shown that
water-table aquifers harbor appreciable numbers of
metabolically active microorganisms, and that these
microorganisms frequently can degrade organic
contaminants. It became necessary to consider
biotransformation in transport models. Unfortunately,
many ground-water scientists adopted the
conceptual model most frequently used to describe
biotransformations in surface waters.
The presence of the contaminant was assumed to
have no effect on microorganism populations that
degrade it. It was also assumed that contaminant
concentration does not influence transformation
kinetics, and that the capacity to transform the
contaminant is uniformly distributed throughout the
body of water under study. These assumptions are
often appropriate for surface waters: contaminant
concentration is usually too low and the residence
time too short to allow adaption of the microbial
community to the contaminant, and the organisms
that are naturally pre-adapted to the contaminant are
mixed throughout the water body by turbulence.
Consequently, utilization kinetics can conveniently be
described by simple first-order decay constants. In
surface waters these constants are usually obtained
by monitoring contaminant disappearance in water
samples.
8.2.3.2 Ground-water biotransformations
These circumstances rarely apply to biotransformation
in ground water. Contaminant residence time is
usually long, at least weeks or months, and frequently
years or decades. Further, contaminant
concentrations that are high enough to be of
environmental concern are often high enough to elicit
adaption of the microbial community. For example,
the U.S. Environmental Protection Agency's
Maximum Contaminant Level (MCL) for benzene is 5
ug/L. This is very close to the concentration of
alkylbenzenes required to elicity adaption to this class
of organic compounds in soils. As a result, the
biotransformation rate of a contaminant in the
subsurface environment is not a constant, but
increases after exposure to the contaminant in an
unpredictable way. Careful field work has shown that
the transformation rate in aquifers of typical organic
contaminants, such as alkylben/enes, can vary as
much as two orders of magnitude over a meter
vertically and a few meters horizontally. This
surprising variability in transformation rate is not
related in any simple way to system geology or
hydrology.
It is difficult to determine first-order rate constants in
subsurface material. Most microbes in subsurface
material are firmly attached to solid surfaces; usually
less than one percent of the total population is truly
planktonic. As a result, the microbes in a ground-
water sample grossly underrepresent the total
microbial population in the aquifer. Thus, contaminant
disappearance kinetics in a ground-water sample do
not represent the behavior of ihe material in the
aquifer. It is therefore necessary to do microcosm
studies with samples representative of the entire
aquifer system - a formidable technical challenge.
156
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8.2.3.3 A Ground-Water Model
These concerns have prompted re-examination of
assumptions about biotransformation implicit or
explicit in a particular modeling approach, with the
realization that no one qualitative description of
biotransformation can be universally applicable. Field
experience has shown that the relationships that
describe the biological fate of contaminants actually
change within aquifers in response to geochemical
constraints on microbial physiology. Rather than
describing biotransformation with a continuous
function applicable at all points in the aquifer, it may
be more realistic to examine key geochemical
parameters and to use that information to identify the
relationship for biotransformation that applies at any
particular point. These key parameters could include
the contaminant concentration, oxygen or other
electron acceptor concentration, redox state, pH,
toxicity of the contaminant or co-occurring materials,
and temperature. One such model has been
evaluated in the field.
The model described an alkylbenzene and
polynuclear aromatic hydrocarbon plume in a shallow
water-table aquifer. Microcosm studies showed that
organisms in the aquifer had adapted to these
contaminants, and would degrade them very rapidly
when oxygen was available. As a result of this
adaption, the hydrocarbon biodegradation rate was
not controlled by any inherent property of the
organisms. Rather, physical transport processes such
as diffusion and dispersion seemed to dominate by
controlling oxygen availability to the plume.
Because the biotransformation rate was controlled by
physical processes, the actual model was very
simple. Oxygen and hydrocarbon transport were
simulated as conservative solutes using the U.S.
Geological Survey method-of-characteristics code.
A subroutine examined oxygen and hydrocarbon
concentrations at each node and generated new
concentrations based on oxidative metabolism
stoichiometry. When the model was projected forward
in time it illustrated an important property of many
such plumes. The plume grew with time until the rate
of admixture of oxygen balanced the rate of release of
hydrocarbons from the source. Afterward, the extent
of the plume was at steady-state.
The body of field experience which can be drawn
upon to properly assign laws for biotransformation is
growing rapidly. Transport-limited kinetics may
commonly apply to releases of petroleum
hydrocarbons and other easily degradable materials
such as ethanol or acetone in oxygenated ground
water. On the other hand, materials that can support
a fermentation, in which an exogenous electron
acceptor is not required, may follow first-order
kinetics. Unfortunately, many important
biotransformations in ground water are still mysteries.
The reductive dehalogenation of small halogenated
hydrocarbons such as trichloroethene and 1,1,1-
trichloroethane is a good example. In such cases
transformation kinetics of the compound are
controlled by transformation kinetics of a second
compound, the primary substrate that supports the
metabolism of the active microorganisms. These
complex interactions are poorly understood and
cannot be described quantitatively at the present
time. However, this is an area of active research, and
hopefully the appropriate relationships may soon be
determined.
Rapid field methods to determine if adaption has
occurred at a site are needed. Tools to predict
whether adaption can be expected, and to estimate
the time required for adaption if it does occur, are
also needed. For systems that are limited by transport
processes, field methods to estimate the aquifer
processes that control mixing, such as transverse
dispersion and exchange processes across the water
table, are required. For systems that are limited by
the intrinsic biotransformation rate, new laboratory
test methods (possibly, improved microcosms) that
will provide reliable estimates of the kinetic
parameters are required.
In addition to being sufficiently accurate and precise,
these new methods should provide estimates that are
truly representative of the hydrologic unit being
simulated. Because contaminants typically have long
residence times in aquifers, slow transformation rates
can have environmental significance. The test
methods should therefore be sufficiently sensitive to
measure transformation rates that are significant in
the hydrologic context being simulated. Finally, there
is a need for models that go beyond simple prediction
of contaminant concentrations at points in the aquifer,
and forecast the concentrations produced by
production wells.
8.2.4 Analytical and Numerical Models
One of the more subtly involved decisions which must
be made is whether to use an analytical model or a
numerical model to solve a particular problem.
Analytical models provide exact solutions, but many
simplifying assumptions must be made for the
solutions to be tractable; this places a burden on the
user to test and justify the underlying assumptions
and simplifications (Javendel et a/., 1984). Fot
example, the Theis equation is an analytical
expression which is used to predict the piezometric
head changes for pumping or injection wells in
confined aquifers (Freeze and Cherry, 1979; Todd,
1980):
s = [Q/(4nT)] x [-0.5772 - ln(u) + u
- (u2/(2 x 2!)) + (u3/(3 x 3!))
- (u4/(4 x 4!)) ...]
where:
s = the change in piezometric head
157
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Q = the flowrate of the well
T = the transmissivity of the aquifer
u = (r2 S) / (4 T t);
r = the radial distance from the well
S = the storage coefficient of the aquifer
t = the length of time the well has been
operating.
Here the principal assumptions are (Lohman, 1972):
o The aquifer is homogenous and isotropic
o The aquifer is of infinite area! extent, relative to
the effects of the well (no boundaries)
o The well is screened over the entire saturated
thickness of the aquifer
o The saturated thickness of the aquifer does not
vary as a result of the operation of the well
o The well has an infinitesimal diameter so that
waters in storage in the casing represent an
insignificant volume
o Water is removed from or injected into the aquifer
with an instantaneous change in the piezometric
head.
Evaluation of the infinite Taylor series representing
the well function integral can be accomplished
graphically using type curves (Walton, 1962; Lohman,
1972). Alternatively, a simplification can be made so
that the Theis equation is directly solvable (Cooper
and Jacob, 1946). This is done by dropping all terms
in the Taylor series with powers greater than one, and
is strictly valid for cases where "u" has a value less
than 0.01 (e.g., Figure 8-3). Physically, this
corresponds to a limitation on the predictive power of
the modified Theis equation; head changes predicted
at locations far from the well are inaccurate, except
for long durations of pumpage (i.e., approaching
equilibrium or steady-state conditions).
Numerical models are much less burdened by these
assumptions and are therefore inherently capable of
addressing more complicated problems, but they
require significantly more data and their solutions are
inexact (numerical approximations). For example, the
assumptions of homogeneity and isotropicity are
unnecessary due to the ability to assign point (nodal)
values of transmissivity and storage. Likewise, the
capacity to incorporate complex boundary conditions
obviates the need for the "infinite areal extent"
assumption. There are, however, difficult choices
facing the user of numerical models; i.e. time steps,
spatial grid designs, and ways to avoid truncation
errors and numerical oscillations must be chosen
(Remson et a/., 1971; Javendel ef a/., 1984). These
choices, if improperly made, may result in errors
unlikely to be made with analytical approaches (e.g.,
mass imbalances, incorrect velocity distributions, and
grid-orientation effects).
8.2.5 Quality Control
These latter points signify a greater need for quality
control measures when contemplating the use of
numerical models. Three levels of quality control have
been suggested previously (Huyakorn ef a/., 1984):
1) Validation of the model's mathematics by
comparison of its output with known analytical
solutions to specific problems,
2) Verification of the general framework of the model
by successful simulation of observed field data,
and
3) Benchmarking of the model's efficiency in solving
problems by comparison with other models.
These levels of quality control address the
soundness and utility of the model alone, and do
not treat questions of its application to a specific
problem. Hence, at least two additional levels of
quality control appear justified:
4) Critical review of the problem conceptualization to
ensure that the modeling effort considers all
physical, chemical, and biological processes
which may affect the problem, and
5) Evaluation of the specifics of the application; e.g.,
appropriateness of the boundary conditions, grid
design, time steps, etc.
Validation of the mathematical framework of a
numerical model is deceptively simple. The usual
approach for ground-water flow models involves a
comparison of drawdowns predicted by the Theis
analytic solution to those obtained by using the
model, such as depicted in Figure 8-4. The
"deceptive" part is the foreknowledge that the Theis
solution can treat only a very simplified situation as
compared with the scope of situations addressable by
the numerical model. In other words, analytical
solutions cannot test most of the capabilities of the
numerical model in a meaningful way; this is
particularly true with regard to simulation of complex
aquifer boundaries and irregular chemical
distributions.
Field verification of a numerical model consists of first
calibrating the model using one set of historical
records (e.g., pumping rates and water levels from a
certain year), and then attempting to predict the next
set of historical records. In the calibration phase, the
aquifer coefficients and other model parameters are
adjusted to achieve the best match between model
outputs and known data; in the predictive phase, no
adjustments are made (excepting actual changes in
pumping rates, etc.). Presuming that the aquifer
coefficients and other parameters were known with
sufficient accuracy, a mismatch means that either the
model is not correctly formulated or that it does not
treat all of the important phenomena affecting the
situation being simulated (e.g., does not allow for
158
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Figure 8-3 Example of plots prepared with the Jacob's
approximation of the Theis analytical solution
to well hydraulics in an artesian aquifer.
Flow rate = 100,000 cu ft /day
Transmissivity = 10,000 sq ft /day
Storage coefficient = 0.0001
2 -
200 400 600 800
Radius of Observation (feet)
1000
Figure 8-4 Mathematical validation of a numerical method
of estimating drawdown, by comparison with
an analytical solution.
2.0
1.8
1.6
1.4
J 1.2
-8 1.0
° .8
.6
.4
.2
Flow rate = 100,000 cu ft /day
Transmissivity = 10,000 sq ft /day
Storage coefficient = 0.00003
Observation radius = 2000 ft
Analytical
(Jacob's approx.)
Numerical
(alt. direct, implicit)
23456789 10
Duration of Pumpage (days)
leakage between two aquifers when this is actually
occurring).
Field verification exercises usually lead to additional
data gathering efforts, because existing data for the
calibration procedure are often insufficient to provide
unique estimates of key parameters. This means that
a "black box" solution may be obtained, which may
be good only for the records used in the calibration.
For this reason, the blind prediction phase is an
essential check on the uniqueness of the parameter
values used. In this regard, field verification of models
using datasets from controlled research experiments
may be much more achievable practically.
Benchmarking routines to compare the efficiency of
different models in solving the same problem have
only recently become available (Ross ef a/., 1982;
Huyakorn ef a/., 1984). Much more needs to be done
in this area, because some unfair perceptions
continue to persist regarding the ostensibly greater
utility of certain modeling techniques. For example, it
has been said many times that finite element models
(FEMs) have an inherent advantage over finite
difference models (FDMs) in terms of the ability to
incorporate irregular boundaries (Mercer and Faust,
1981); the number of points (nodes) which must be
used by FEMs is considerably less due to the flexible
nodal spacings that are allowed. Benchmarking
routines, however, show that the much longer
computer time required to evaluate FEM nodes
causes there to be little, if any, cost advantage for
simulations of comparable accuracy.
8.3 Applications in Practical Settings
8.3.1 Stereotypical Applications
As stated in preceding sections, models are
simplifications of reality that may or may not faithfully
simulate the actual situation. Typically, attempts are
made to mimic the effects of hydrogeologic,
chemical, and biological processes in practical
applications of models. These almost always involve
idealizations of known or suspected features of the
problem on hand. For example, the stratification of
alluvial, fluvial, and glacial deposits may be assumed
to occur in uniformly thick layers, despite the great
variability of stratum thicknesses found in actual
settings. Large blocks of each stratum are assumed
to be homogenous. Sources of chemical input are
commonly assumed to have released contaminants at
constant rates over the seasons and years of
operational changes that the sources were active.
The areal distribution of rainfall and the actual
schedules of pumpage from production wells are also
artificially homogenized in most modeling exercises.
All these idealizations are made necessary by a lack
of the appropriate historical records and field-derived
parameter estimates, and all reduce the reliability of
predictions made with models. The degree of
159
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usefulness of a model is therefore directly dependent
on the subjective judgments that must be made in
data collection and preparation efforts prior to
attempting mathematical simulations. This is true not
only in a quantitative sense, but also in a qualitative
sense because it is the data gathering phase of a
project that begets the conceptualization on which the
model will be based.
8.3.2 Real-World Applications
To illustrate this point, the highlights of two very
different contamination problems will be described.
The first involves a relatively limited contamination
incident arising from a very small source and having
few contaminants. The second involves a major
contamination incident arising from the operation of a
chemical reprocessing facility that handled dozens of
different contaminants in large amounts. The common
theme that is shared by the two cases, as should also
apply to virtually all cases, is one of seeking to define
the relative influences of natural processes affecting
contaminant transport in order to optimize the
assessment and remediation of the problem. It is the
validity of the conceptual model of what is happening
at these sites that is most important, not the
application of a particular mathematical model.
8.3.2.1 Field example no. 1
The Lakewood Water District in Lakewood,
Washington, operates a number of wells for drinking
water supply purposes. Some of the wells operated
by the District, such as the two primary wells at its
Render's Corner site (Figures 8-5, 8-6), have been
contaminated (USEPA-Region 10, 1981) with low
levels of volatile organic chemicals (VOCs). During
the course of the investigations at the Render's
Corner site, a number of cost-saving sampling
alternatives were chosen. These related principally to
the field use of a portable gas chromatograph
(Organic Vapor Analyzer) for the screening of water
samples and soil extracts taken while drilling
monitoring wells, and to the use of selective analyses
(volatiles only) of ground-water samples when initial
results showed only a narrow group of contaminants
to be present. The lowered analytical costs, in part,
allowed for increased expenditures for geotechnical
characterization of the site (Wolf and Boateng, 1983).
The geotechnical efforts, particularly the pump tests
which were conducted, led to a realization that the
source of the contaminants was to be found
regionally downgradient (Keely and Wolf, 1983). The
pumping strength of the water-supply wells, when
operating, was sufficient to pull contaminants over
400 feet back against the regional flow direction.
Because most contaminant sources are found
upgradient of the wells they affect, this behavior was
somewhat unexpected. A unique feature of the field
investigation was the taking of ground-water
samples from the pumping wells concurrent with
drawdown measurements obtained during pump tests
(Keely, 1982).
Figure 8-5 Location map for Lakewood Water District
wells contaminated with volatile organic
chemicals.
Lakewood Study
Well Location Map
Not Drawn to Scale
13
12
The pump tests yielded estimates of local
transmissivity and storage coefficients. It also
confirmed the presence of a major aquifer boundary
nearby; a buried glacial till drumlin just west of the
site parallels the general direction (north) of regional
flow. The pump tests clearly showed some anisotropy
of the sediments as well; drawdown contours
produced an elliptical cone of drawdown, the major
axis of which was aligned with the regional flow to the
north. This information resulted in modifications to the
original plans, which called for drilling and
constructing several monitoring wells west of the site.
Instead, more monitoring wells were drilled along the
north-south axis. Chemical analysis of the samples
taken concurrent with drawdown measurements
formed a time-series of contaminant concentrations
that provided a clue to where the contaminant source
was located (Keely, 1982). The time-series showed
that the well nearest the downgradient edge of the
well field was exposed to increasing contaminant
levels as pumping continued, whereas the upgradient
pumping well remained largely unaffected (Keely and
Wolf, 1983).
The hydrogeologic parameter estimates obtained from
the pump tests strengthened the conceptualization of
contaminants being drawn back against the regional
flow because the capture zones of the pumping wells
were sufficiently distorted by the local anisotropy to
more than encompass the contaminant source.
Without considering the anisotropic bias along the
regional flow path, the estimated boundaries of the
capture zone for either of the two wells marginally
reached the distance to the contaminant source. The
160
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Figure 8-6 Geologic logs for Lakewood Water District
wells contaminated with volatile organic
chemicals.
Elevation Well
MSL H-2
(1959)
Well
H-1
(feet) (meters)
275-i
-80
250-
(1951)
-70
225-
200-
175-
-60
.«•»••
Gravel &
Sand, Loose
Hardpan
(Till)
Gravel &
Sand, Tight
Some Water
below 275 ft
Gravel &
Sand, Water
Bearing
Gravel,
Dirty
Hardpan
w/ Large
Boulders
Sand ft
Gravel
Hardpan
Sand &
Gravel,
Dirty
Gravel,
Loose
-50
mechanism by which the two wells became
contaminated seemed to be understood from a
hydraulic point of view (Figure 8-7), but the chemical
information did not seem to provide a consistent
picture.
The source of contamination, a septic tank at a dry-
cleaning facility, was found to have received large
amounts of tetrachloroethylene and trichloroethylene,
but no known amounts of cis- or trans-
dichloroethylene; whereas the contaminated wells had
relatively high concentrations of dichloroethylene.
Initially it was thought that other sources might also
be present and would explain the high concentrations
of dichloroethylene. However, it soon became clear
that recent research results regarding the potential for
biotransformation of tetrachloroethylene and
trichloroethylene (Wilson and McNabb, 1981) would
more satisfactorily explain the observations.
Simulations of this kind of problem could be
adequately performed only by contaminant transport
models capable of incorporating the effects of the
pumping wells on the regional flow field. More
sophisticated approximations would also require the
ability to account for the anisotropic and
Figure 8-7 Schematic illustrating the mechanism by which
a downgradient source may contaminate a
production well, and by which a second well
may isolate the source through hydraulic
interference.
Well Well
A. H-2 H-1
Water Table
*~ n** J »
<=
B.
Q: 1175gpm
Water Table
Q: 875 gpm
hexterogeneous character of the site, the retardation
of the VOCs by sorption, and their possible
biotransformations. Given the higly localized nature of
the contaminant source and limited extent of the
plume, however, there was insufficient justification for
pursuing such efforts. The resolution of the problem
was possible by relatively simple source removal
techniques (excavation of the septic tank and
elimination of discharges).
8.3.2.2 Field example no. 2.
Similar experience with special use of geotechnical
methods and state-of-the-art research findings
occurred at the 20-acre Chem-Dyne solvent
reprocessing site in Hamilton, Ohio (Figure 8-8).
During operation of the site (1974-1980), poor waste
handling practices such as on-site spillage of a wide
variety of industrial chemicals and solvents, direct
discharge of liquid wastes to a stormwater drain
beneath the site, and mixing of imcompatible wastes
were engaged in routinely. These caused extensive
soil and ground-water contamination, massive fish
kills in the Great Miami River, and major on-site fires
and explosions, respectively. The stockpiling of liquid
and solid wastes resulted in thousands of badly
161
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Figure 8-8 Location map for Chem-Dyne Superfund Site.
ICHEM-DYNE SUPERFUND SITE
Hamilton, Ohio
LEGEND
• monitoring well locations
MW1 monitoring well identity
" • • site boundary
LOCATION MAP
corroded leaking drums that posed a long-term
threat to the environment (Ch^M-Hill, 1984).
The seriousness of the ground-water contamination
problem became evident during the initial site survey
(1980-1981), which included the construction and
sampling of over 20 shallow monitoring wells (Ecology
and Environment, 1981). The initial survey indicated
that the contaminant problem was much more limited
than was later shown to be the case (Roy F. Weston
Inc., 1983, Ch2M-Hill, 1984). A good portion of the
improvement in delineating the plume was brought
about by an improved understanding of the natural
processes controlling transport of contaminants at the
site.
The initial site survey indicated that ground-water
flow was generally to the west of the site, toward the
Great Miami River, but that a shallow trough
paralleled the river itself as a result of weak and
temporary stream influences. The study concluded
that contaminants would be discharged from the
aquifer into the river (Ecology and Environment,
1981). That study also concluded that the source was
limited to highly contaminated surface soils, and that
removal of the uppermost three feet of the soil would
essentially eliminate the source.
That conclusion was, however, based on faulty soil
sampling procedures. The soil samples that were
taken were not preserved in air-tight containers, so
that most of the VOCs leaked out prior to analysis.
That the uppermost soil samples showed high VOC
levels is probably explained by the co-occurrence of
viscous oils and other organic chemicals that may
have served to entrap the VOCs. The more fiscous
and highly retarded chemicals did not migrate far
enough into the vertical profile 1o exert a similar
influence on samples collected at depths greater than
a few feet.
Subsequent studies of the site corrected these
misinterpretations by producing data from proper soil
samplings and by incorporating much more detailed
characterization of the fluvial sediments and the
natural flow system. In those studies vertical profile
characterizations were obtained from each new
borehole drilled by continuous split-spoon samples
of subsurface solids; and clusters of vertically-
separated monitoring wells were constructed. The
split-spoon samples helped to confirm the general
locations of intefingered clay lenses and clearly
showed the high degree of heterogeneity of the
sediments (Figures 8-9 and 8-10). While an
extensive network of shallow wells confirmed earlier
162
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indications of general ground-water flow toward the
river (Figure 8-11), the clusters of vertically
separated wells revealed that dramatic downward
gradients existed adjacent to the Great Miami River
(Figure 8-12). This finding indicated that the
migrating plume would not be discharged to the river,
but would instead flow under the river.
The presence of major industrial wells on the other
side of the river supported this conclusion. The plume
would be drawn to greater depths in the aquifer by
the locally severe downward gradient, but whether the
industrial wells would actually capture the plume
could not be determined. That determination would
require careful evaluation of the hydrogeologic
features beneath the river; something that has not
been attempted because of the onset of remedial
actions designed to stop the plume from reaching the
river.
The field characterization efforts, however, did include
the performance of a major pump test so that the
hydrogeologic characteristics of the contaminated
portion of the aquifer could be estimated. The pump
test was difficult to arrange, because the pumping
well had to be drilled onsite for reasons of potential
liability and lack of property access elsewhere. The
drillers were considerably slowed in their work by the
need to don air-tanks when particularly
contaminated subsoils were encountered because the
emission of volatile fumes from the borehole
presented unacceptable health risks. Since the waters
which would be pumped were expected to be
contaminated, it was necessary to construct 10 large
temporary holding tanks (100,000 gallons each)
onsite to impound the waters for testing and possible
treatment prior to being discharged to the local sewer
system (CH2M-HJII, 1984).
The costs and difficulty of preparing for and
conducting the test were worth the effort, however.
The water levels in thirty-six monitoring wells were
observed during the test and yielded a very detailed
picture of transmissivity variations (Figure 8-13),
which has been used to help explain the unusual
configuration of the plume (Figure 8-14) and which
were used to guide the design of a pump-and-treat
system. Storage coefficients were also estimated; and
though the short duration of the test (14 hours) did
not allow for definitive estimates to be obtained, it
was clear that qualitative confirmation of the generally
non-artesian (water-table) nature of the aquifer
beneath the site was confirmed. An automated data
acquisition system (computer controlled pressure
transducer) was used to monitor the water levels and
provide real-time drawdown plots of 19 of the 36
wells (Table 8-2), greatly enhancing the information
obtained with only minimal manpower requirements.
The benefits from conducting the pump test cannot
be overemphasized; qualitative confirmation of
lithologic information and semi-quantitative
estimation of crucial parameters were obtained.
Finally, the distribution patterns of contaminant
species that emerged from the investigations at
Chem-Dyne were made understandable by
considering research results and theories regarding
chemical and microbiological influences. Once again
there seemed to be evidence of transformation of
tetrachloroethene (Figure 8-15) to less halogenated
daughter products such as trichloroethene (Figure 8-
16), dichloroethene (Figure 8-17), and vinyl
chloride/monochloroethene (Figure 8-18). The
relative rates of movement of these contaminants, as
well as other common solvents like benzene (Figure
8-19) and chloroform (Figure 8-20), generally
conformed to predictions based on sorption
principles. The remediation efforts also made use of
these contaminant transport theories in estimating the
capacity of the treatment system needed and the
length of time necessary to remove residuals from the
aquifer solids (CH2M-Hill, 1984).
During the latter stages of negotiations with the
Potentially Responsible Parties (PRPs), government
contractors prepared mathematical models of the flow
system and contaminant transport at Chem-Dyne
(GeoTrans, 1984). These were used to estimate the
possible direction and rate of migration of the plume
in the absence of remediation, the mass of
contaminants removed during various remedial
options, and the effects of sorption and dispersion on
those estimates. Because of the wide range of
sorption properties associated with the variety of
VOCs found in significant concentrations it was
necessary to select values of retardation constants
that represented the likely upper- and lower-limits
of sorptive effects. It was also necessary to estimate
or assume the values of other parameters known to
affect transport processes, such as dispersion
coefficients.
While the developers of the models would be the first
to acknowledge the large uncertainties associated
with those modeling efforts due to lack of information
about the actual history of chemical inputs and other
important data, there was agreement between the
government and PRP technical experts that the
modeling efforts had been very helpful in assessing
the magnitude of the problem and in determining
minimal requirements for remediation. Consequently,
modeling efforts will continue at Chem-Dyne. Data
generated during the remediation phase will be used
to refine models in an ongoing process so that the
effectiveness of the remedial action can be evaluated
properly.
8.3.3 Practical Concerns
In many ways, there may be too much confidence
among those not directly involved in ground-water
quality research regarding current abilities to predict
163
-------
Figure 8-9 Chem-Dyne geologic cross-section along NNW-SSE axis.
NNW
600
590
580
570
560
* 550
c
o
1
1 540
530
520
510
500
SSE
MW14
MW28
Water Level Elevation
Approximately 563 ft MSL
October 30, 1983
MW26
MW10
p. ?.,'.<
MW24
MW9
0
,.p-
MW8
DATA SOURCE: U.S. EPA, 1984
Fill (sandy gravel)
Clayey silt, silty clay
Sandy gravel, gr. sand
Silty sand
Clayey gravel, glacial till
164
-------
Figure 8-10 Chem-Dyne geologic cross-section along WSW-ENE axis.
wsw
ENE
R.QO
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Fill (sandy gravel)
Sand
Siltysand
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Sandy gravel, gr. sand
Clayey gravel, glacial till
165
-------
Figure 8-11 Shallow well ground-water contour map for Chem-Dyne. Flow is generally to the river (west) and down the
valley (southwest).
HYDRAULIC CANAL
LEGEND
• monitoring well location
MW1 monitoring well identity
- • • site boundary
,x*^-x water level elevation contour
CHEM-DYNE SUPERF:UND SITE
Hamilton, Ohio
April, 1983 Water Level Elevation Contours
for Shallow Wells
(values in feet)
166
-------
Figure 8-12 Typical arrangement of clustered, vertically-
separated wells installed adjacent to Chem-
Dyne and the Great Miami River.
Ground
Surface
Shallow
Well
Intermediate
Well
Deep
Well
>
transport and fate of contaminants in the subsurface.
The discussions in the preceding sections should
place in proper perspective the admittedly remarkable
advances that have been made in recent years by
illustrating the practical and conceptual uncertainties
that remain unresolved. Continuing research efforts
will eventually resolve these uncertainties, but those
efforts will be considerably slower if existing results
are not routinely incorporated into practical situations.
Research results must be tested in real-world
settings because there is no alternative mechanism
for validating them. Just as importantly, there are
economic arguments for incorporating research
findings and state-of-the-art techniques into
routine contaminant investigations and remediations.
Additional effort devoted to site-specific
characterizations of natural process parameters,
rather than relying almost exclusively on chemical
analyses of ground-water samples, can significantly
improve the quality and cost-effectiveness of
remedial actions at such sites. To underscore this
point, condensed summaries are provided of the
principal activities, benefits, and shortcomings of
three possible site characterization approaches:
conventional (Table 8-3), state-of-the-art (Table
8-4), and state-of-the-science (Table 8-5). To
further illustrate this, a qualitative assessment of
desired trade-offs between characterization and
clean-up costs is presented in Figure 8-21.
As illustrated there, some investments in specialized
equipment and personnel will be necessary to make
transitions to more sophisticated approaches, but
those investments should be more than paid back in
reduced clean-up costs. The maximum return on
increased investments is expected for the state-of-
the-art approach, and will diminish as the state-
of-the-science approach is reached because highly
specialized equipment and personnel are not widely
available. It is vitally important that this philosophy be
considered, because the probable benefits in lowered
total costs, health risks, and time for effective
remediations can be substantial.
8.4 Liabilities, Costs, and
Recommendations for Managers
There are many texts available that describe the
derivation of the theories underlying mathematical
models, the technical applications of models, and
related technical topics (e.g., data collection and
parameter estimation techniques). Few texts treat the
nontechnical issues that managers face when
evaluating the possible uses of models, such as
potential liabilities, costs, and communications
between the modeler and management. These are,
however, important considerations because many
modeling efforts fail as a consequence of insufficient
attention to them. This section is therefore directed to
those issues.
8.4.7 Potential Liabilities
Some of the liabilities attending the use of
mathematical models relate to the degree to which
predictive models are relied on to set conditions for
permitting or banning specific practices or products. If
a model is incapable of treating specific applications
properly, substantially incorrect decisions may be
made. Depending on the application, unacceptable
environmental effects may begin to accumulate long
before the nature of the problem is recognized.
Conversely, unjustified restrictions may be imposed
on the regulated community. Inappropriate or
inadequate models may also cause the 're-opening
clause' of a negotiated settlement agreement to be
invoked when, for instance, compliance requirements
that were guided by model predictions of expected
plume behavior are not met.
Certain liabilities relate to the use of proprietary codes
in legal settings, where the inner workings of a model
may be subject to disclosure in the interests of
justice. The desire for confidentiality by the model
developer would likely be subordinate to the public
right to full information regarding actions predicated
on modeling results. The mechanisms for protection
of proprietary rights do not currently extend beyond
extracted promises of confidentiality by reviewers or
167
-------
Figure 8-13 Estimates of transmissivity obtained from shallow and deep wells during Chem-Dyne pump lest.
LEGEND
• monitoring well locations
MW1 monitoring well identity
• • • site boundary
CHEM-DYNE SUPERFUND SITE
Hamilton, Ohio
Transmissivity Estimate from October
1983 Pump Test
(values in thousands of square feet per day)
168
-------
Figure 8-14 Distribution of total volatile organic chemical contamination in shallow wells at Chem-Dyne during October,
1983 sampling.
HYDRAULIC CANAL
LEGEND
• monitoring well locations
MW1 monitoring well identity
• - - site boundary
-'^^ Isopleth in parts per billion
(shallow wells only)
CHEM-DYNE SUPERFUND SITE
Hamilton, Ohio
TOTAL VOLATILE ORGANIC CHEMICALS
OCTOBER 1983 SAMPLING
169
-------
Table 8-2 Chem-Dyne Pump Test Observation Network
Observation
Well Number
MW1
MW2
MW3
MW4
MW5
MW6
MW7
MW8
MW9
MW10
MW11
MW12
MW13
MW14
MW15
MW16
MW17
MW18
MW19
MW20
MW21
MW22
MW23
MW24
MW25
MW26
MW27
MW28
MW29
MW30
MW31
MW32
MW33
MW34
MW35
MW36
Pumping
Well
Radial
Distance
(ft)
957
965
848
537
313
420
480
740
487
186
502
232
232
701
611
1275
1518
692
1204
1225
1259
1261
298
398
53
62
272
248
167
993
465
1236
690
454
651
696
0
(reference point)
Initial
Water Level
(ft, MSL)
563.68
563.74
563.96
563.27
563.31
564.40
563.30
563.01
563.08
563.29
562.90
562.39
563.19
—
563.10
562.47
560.03
562.67
559.80
562.10
561.29
559.95
563.07
563.07
563.04
562.96
563.13
562.99
563.23
561.25
562.78
559.56
562.06
562.29
562.93
562.69
562.97
Method of
Measurement
(type, field unit)
Manual, electric probe
Manual, electric probe
Automatic, float-type
Manual, electric probe
Manual, electric probe
Manual, electric probe
Manual, electric probe
Manual, electric probe
Manual, electric probe
Automatic, pressure transducer
Automatic, pressure transducer
Automatic, pressure transducer
Automatic, pressure transducer
Dry — no data collected
Automatic, pressure transducer
Manual, electric probe
Manual, electric probe
Manual, electric probe
Automatic, float-type
Automatic, float-type
Manual, electric probe
Automatic, pressure transducer
Automatic, pressure transducer
Automatic, pressure transducer
Automatic, pressure transducer
Automatic, pressure transducer
Automatic, pressure transducer
Automatic, pressure transducer
Automatic, pressure transducer
Manual, electric probe
Automatic, float-type
Automatic, pressure transducer
Automatic, pressure transducer
Automatic, pressure transducer
Automatic, pressure transducer
Manual, electric probe
Automatic, pressure transducer
other interested parties. Hence, a developer of
proprietary codes still assumes some risk of exposure
of innovative techniques, even if the code is not
pirated outright.
Yet other liabilities may arise as the result of
misapplication of models or the application of models
later found to be faulty. Frequently, the choices of
boundary and initial conditions for a given application
are hotly contested; misapplications of this kind are
undoubtedly responsible for many of the reservations
expressed by would-be model users. It has also
happened many times in the past that a widely used
and highly regarded model code was found to contain
errors that affected its ability to faithfully simulate
situations for which it was designed. The best way to
minimize these liabilities is to adopt strict quality
control procedures for each application.
8.4.2 Economic Considerations
The nominal costs of the support staff, computing
facilities, and specialized graphics' production
equipment associated with numerical modeling efforts
can be high. In addition, quality control activities can
result in substantial costs; the determining factor in
controlling these costs is the degree to which a
manager must be certain of the characteristics of the
model and the accuracy of its output.
As a general rule, costs are greatest for personnel,
moderate for hardware, and minimal for software. The
exception to this ordering relates to the combination
of software and hardware purchased. An optimally
outfitted business computer (e.g., VAX 11/785 or IBM
3031) costs about $100,000; but it can rapidly pay for
itself in terms of dramatically increased speed and
computational power. A well complimented personal
computer (e.g., IBM-PC/AT or DEC Rainbow) may
cost $10,000; but the significantly slower speed and
170
-------
Figure 8-15 Distribution of tetrachloroethane in shallow wells at Chem-Dyne during October, 1983 sampling.
LEGEND
• monitonng well locations
MW1 monitoring well identity
. . . site boundary
Isopleth in parts per billion
(shallow wells only)
CHEM-DYNE SUPERFUND SITE
Hamilton, Ohio
TETRACHLOROETHENE
OCTOBER 1983 SAMPLING
171
-------
Figure 8-16 Distribution of trichloroethane in shallow wells at Chem-Oyne during October, 1983 sampling.
HYDRAULIC CANAL
LEGEND
• monitoring well locations
MW1 monitoring well identity
. . . site boundary
Isopleth in parts per billion
(shallow wells only)
CHEM-DYNE SUPERFUND SITE
Hamilton, Ohio
TRICHLOROETHENE
OCTOBER 1983 SAMPLING
172
-------
Figure 8-17 Distribution of trans-dichloroethene in shallow wells at Chem-Dyne during October, 1983 sampling.
HYDRAULIC CANAL
300 1 1
lOOm
1
LEGEND
• monitoring well locations
MW1 monitoring well identity
- . - site boundary
-**^* Isopleth in parts per billion
(shallow wells only)
CHEM-DYNE SUPERFUND SITE
Hamilton, Ohio
trans-DICHLOROETHENE
OCTOBER 1983 SAMPLING
173
-------
Figure 8-18 Distribution of vinyl chloride in shallow wells at Chem-Oyne during October, 1983 sampling.
LEGEND
• monitoring well locations
MW1 monitoring well identity
... site boundary
Isopleth in parts per billion
(shallow wens only)
CHEM-DYNE SUPERFUND SITE
Hamilton, Ohio
VINYL CHLORIDE
(MONOCHLOROETHENE)
OCTOBER 1983 SAMPLING
174
-------
Figure 8-19 Distribution of benzene in shallow wells at Chem-Dyne during October, 1983 sampling.
LEGEND
• monitoring well locations
MW1 monitoring well identity
. • . site boundary
Isopleth in parts per billion
(shallow wells only)
CHEM-DYNE SUPERFUND SITE
Hamilton, Ohio
BENZENE
OCTOBER 1983 SAMPLING
175
-------
Figure 8-20 Distribution of chloroform in shallow wells at Chem-Dyne during October, 1983 sampling.
LEGEND
• monjtoring well locations
MW1 monitoring well identity
... site boundary
Isopleth in parts per billion
(shallow wells only)
CHEM-DYNE SUPERFUND SITE
Hamilton, Ohio
CHLOROFORM
OCTOBER 1983 SAMPLING
176
-------
Table 8-3 Conventional Approach to Site
Characterization Efforts
Actions Typically Taken
Install a few dozen shallow monitoring wells
Sample and analyze numerous times for 129+ pollutants
Define geology primarily by driller's log and cuttings
Evaluate hydrology with water level maps only
Possibly obtain soil and core samples (chemical extractions)
Benefits
Rapid screening of problem
Moderate costs involved
Field and lab techniques standardized
Data anlaysis relatively straightforward
Tentative identification of remedial options possible
Shortcomings
True extent of problem often misunderstood
Selected remedial alternative may not be appropriate
Optimization of remedial actions not possible
Clean-up costs unpredictable and excessive
Verification of compliance uncertain and difficult
Table 8-5 State-of-the-Science Approach to Site
Characterization Efforts
Idealized Approach
Assume "State-of-the-Art Approach" as starting point
Conducter tracer tests and borehole geophysical surveys
Determine % organic carbon, exchange capacity, etc. of
solids
Measure redox potential, pH, dissolved oxygen, etc. of
fluids
Evaluate soprtion-desorption behavior using select cores
Identify bacteria and assess potential for biotransformation
Benefits
Thorough conceptual understanding of problem obtained
Full optimization of remedial actions possible
Predictability of remediation effectiveness maximized
Clean-up costs lowered significantly; estimates reliable
Verification of compliance assured
Shortcomings
Characterization costs significantly higher
Few previous field applications of advanced theories
Field and laboratory techniques not yet standardized
Availability of specialized equipment low
Demand for specialists dramatically increased
Figure 8-21 General relationship between site
characterization costs and clean-up costs as
a function of the characterization approach.
High
Table 8-4 State-of-the-Art Approach to Site
Characterization Efforts
Recommended Actions
Install depth-specific well clusters
Sample and analyze for 129+ pollutants initially
Analyze selected contaminants in subsequent samplings
Define geology by extensive coring/split-spoon samples
Evaluate hydrology with well clusters and geohydraulic tests
Perform limited tests on solids (grain size, clay content)
Conduct geophysical surveys (resistivity soundings, etc.)
Benefits
Conceptual understanding of problem more complete
Better prospect for optimization of remedial actions
Predictability of remediation effectiveness increased
Clean-up costs lowered; estimates improved
Verification of compliance soundly based, more certain
Shortcomings
Characterization costs somewhat higher
Detailed understanding of problem still difficult
Full optimization of remedial actions not likely
Field tests may create secondary problems
Demand for specialists increased
limited computational power may infer hidden costs in
terms of the inability to perform specific tasks. For
example, highly desirable statistical packages like
SAS and SPSS are unavailable or available only with
reduced capabilities for personal computers; many of
the most sophisticated mathematical models are
available in their fully-capable form only on business
computers.
Figure 8-22 gives a brief comparison of typical costs
for software for different levels of computing power.
Obviously, the software for less capable computers is
S
Low
Conventional
Approach
State-of-
the-Art
State-of-
the-Science
cheaper, but the programs are not equivalent; so
managers need to thoroughly think through what level
is appropriate. If the decisions to be made are to be
based on very little data, it may not make sense to
insist on the most elegant software and hardware. If
the intended use involves substantial amounts of data
and sophisticated analyses are desired, it would be
unwise to opt for the least expensive combination.
Based on experience and observation, there does
seem to be an increasing drive away from both ends
of the spectrum and toward the middle; that is, the
use of powerful personal computers is increasing
rapidly, whereas the use of small programmable
calculators and large business computers alike is
177
-------
Figure 8-22 Average price per category for ground-water
models from the International Ground Water
Modeling Center.
2
£
~o
O
2
<
00
80
60
40
20
n
-
-
-
-
—*
wwwm
wvmt
n
1 1 n
12345
Ground-Water Modeling Software Categories
Categories
1 Mainframe / business computer models
2 Personal computer versions of mainframe models
3 Original IBM-PC and compatibles' models
4 Handheld microcomputer models (e.g., Sharp
PC1500)
5 Programmable calculator models (e.g., HP41-CV)
Prices include software and all available
documentation, reports, etc.
declining. In part, this stems from the significant
improvements in the computing power and quality of
printed outputs obtainable from personal computers.
In part, it is due to the improved telecommunications
capabilities of personal computers, which are now
able to emulate the interactive terminals of large
business computers so that vast computational power
can be accessed and the results retrieved with no
more than a phone call. Most importantly for ground-
water managers, many of the mathematical models
and data packages have been "down-sized" from
mainframe computers to personal computers; many
more are being written directly for this market.
Since it is expected that most managers will want to
explore this situation a bit more, Figure 8-23 has
been prepared to provide some idea of the costs of
available software and hardware for personal
computers.
The technical considerations discussed in previous
sections indicate that the desired accuracy of the
modeling effort directly affects the total costs of
mathematical simulations. Thus managers will want to
determine the incremental benefits gained by
increased expenditures for more involved
mathematical modeling efforts. There are many
economic theories which can be helpful in
determining the incremental benefits gained per
increased level of investment. The most
straightforward of these are the cost-benefit
approaches commonly used to evaluate the economic
desirability of water resource projects. There are two
generalized approaches in common practice: the
"benefit/cost ratio" method and the "net benefit"
method.
The benefit/cost ratio method involves tallying the
economic value of all benefits and dividing that sum
by the total costs involved in generating those
benefits (i.e., B/C = ?). A ratio greater than one is
required for the project to be considered viable,
though there may be sociopolitical reasons for
proceeding with projects that do not meet this
criterion. Consider the example of a project that is
about to get underway and has gained considerable
social or political momentum when the initial cost
estimates begin to prove to be too low. Not
proceeding or substantially altering the work may be
economically wise; however, such a decision may be
viewed as a breach of faith by the public. Regardless
of how this kind of situation evolves, it is not
uncommon for certain costs to be forgiven or
subsidized, which muddies the picture for incremental
benefits or trade-off analyses.
The "net benefit" method involves determining the
arithmetic difference of the total benefits and total
costs (i.e., B-C = ?). Here the obvious criterion is
that the proposed work results in a situation where
total benefits exceed total costs. This approach is
most often adopted by profit-making enterprises,
because they seek to maximize the difference as a
source of income. The ratio method, by contrast, has
long been used by government agencies and other
non-profit organizations because they seek to show
the simple viability of their efforts irrespective of the
costs involved.
In a very real sense, then, these two general
economic assessment methods stem from different
philosophies. They share many common difficulties
and limitations, however. For example, there is a
need to predict the present worth of future costs and
to amortize benefits over the life of a project. The
mechanics of such calculations are well known, but
they necessarily involve substantial uncertainties. For
example, the present worth of a series of equal
payments for equipment or software can be
computed by (White ef a/., 1984):
P = Ax((1+ i)[n]-1)/(ix(1+ i)[n]) (8-2)
where:
P = present worth
A = series payment each interest period
i = interest rate per period
n = number of interest periods.
Note, however, that the interest rate must be
estimated; this has fluctuated widely in the past two
decades as a result of inflationary and recessionary
periods in our economy. The significance of this is
that a small difference in the interest rate results in
178
-------
Figure 8-23 Price ranges for IBM-PC ground-water models available from various sources.
"5
Q
w
5
10
-------
tremendous differences in the present worth estimate
because of the exponential nature of the equation.
It is also possible to compute the future worth of a
present investment, to calculate the percentage of
worth annually acquired through single payments or
serial investments, and so on. One should be aware
that these methods of calculating costs belong to the
general family of "single-objective", or "mutually-
exclusive alternative" analyses which presuppose that
the cost of two actions is obtained by simple addition
of their singly-computed costs. In other words, the
efforts being evaluated are presumed to have no
interactions. For some aspects of ground-water
modeling efforts, this assumption may not be valid;
e.g., one may not be able to specify software and
hardware costs independently. In addition, these
methods rely on the "expected value concept",
wherein the expected value of an alternative is viewed
as the single product of its effects and the probability
of their occurrence. This means that high-risk, low-
probability alternatives and low-risk, high-probability
alternatives have the same expected value.
To overcome these difficulties it is necessary to use
methods which can incorporate functional
dependencies between various alternatives and which
do not rely on the expected value concept, such as
multi-objective decision theories (Asbeck and
Haimes, 1984; Haimes and Hall, 1974; Haimes,
1981). A conceivable use would be the estimation of
lowered health risks associated with various remedial
action alternatives at a hazardous waste site. In such
a case the output of a contaminant transport model
would be used to provide certain inputs (i.e., water
levels, contaminant concentrations, etc.) to a health
effects model, and it would convert these into the
inputs for the multiobjective decision model (e.g.,
probability of additional cancers per level of
contaminant). The primary difficulty with these
approaches to cost-benefit analyses is in clearly
formulating the overall probabilities of the alternatives,
so that the objectives which are to be satisfied may
be ranked in order of importance. A related difficulty
is the need to specify the functional form of the inputs
(e.g., the "population distribution function" of
pumpage rates or contaminant levels). Historical
records about the inputs may be insufficient to allow
their functional forms to be determined.
Another problem compounding the cost-benefit
analysis of mathematical modeling efforts relates to
the need to place an economic value on intangibles.
For example, the increased productivity a manager
might expect as a result of rapid machine calculations
replacing hand calculations may not be as definable in
terms of the improved quality of judgments made as it
is in terms of time released for other duties. Similarly,
the estimation of improved ground-water quality
protection benefits may necessitate some valuation of
the human life and suffering saved (rather nebulous
quantities). Hence, there is often room for
considerable "adjustment" of the values of costs and
benefits. This flexibility can be used inappropriately to
improve otherwise unsatisfactory economic
evaluations. Lehr (1986) offers a scathing indictment
of the Tennessee Valley Authority for what he
described as an "extreme injustice", perpetrated by
TVA in the form of hydroelectric projects which have
"incredibly large costs and "negative cost benefit
ratios".
Finally, some costs and benefits may be incorrectly
evaluated because the data on which they are based
are probabilistic and this goes unrecognized. For
instance, we often know the key parameters affecting
ground-water computations (i.e., hydraulic
conductivity) only to within an order of magnitude due
to data collection limitations. In these situations great
caution must be exercised. On the one hand,
excessive expenditures may be made to ensure that
the model "accurately" simulates observed (though
inadequate) data. On the other, the artistic beauty of
computer generated results sometimes generates its
own sense of what is "right", regardless of apparent
clashes with common sense. The reason the basic
data are uncertain is very important. Costs are not
uncertain just because of lack of information about
future interest rates; many times expectations are not
realized because of societal and technological
changes. Miller (1980) noted that EPA overestimated
the cost of compliance with its proposed standard for
vinyl chloride exposure by 200 times the actual costs.
8.4.3 Managerial Considerations
The return on investments made to use mathematical
models rests principally with the training and
experience of the technical support staff applying the
model to a problem, and on the degree of
communication between those persons and
management. In discussing the potential uses of
computer modeling for ground-water protection
efforts, Faust and others (1981) summarized by
noting that "the final worth of modeling applications
depends on the people who apply the models".
Managers should be aware that a fair degree of
specialized training and experience are necessary to
develop and apply mathematical models, and
relatively few technical support staff can be expected
to have such skills presently (van der Heijde et a/.,
1985). This is due in part to the need for familiarity
with a number of scientific disciplines, so that the
model may be structured to faithfully simulate real-
world problems.
What levels of training and experience are necessary
to apply mathematical models properly? Do we need
"Rennaissance" specialists or can interdisciplinary
teams be effective? The answers to these questions
are not clear-cut. From experience it is easy to see
that the more informed an individual is, the more
effective he or she can be. H is doubtful, however,
180
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that any individual can master each discipline with the
same depth of understanding that specialists in those
fields have. What is clear is that some working
knowledge of many sciences is necessary so that
appropriate questions may be put to specialists, and
so that some sense of integration of the various
disciplines can evolve. In practice this means that
ground-water modelers have a great need to
become involved in continuing education efforts.
Managers should expect and encourage this because
the benefits to be gained are tremendous, and the
costs of not doing so may be equally large.
An ability to communicate effectively with
management is essential also. Just as is the case
with statistical analyses, an ill-posed problem yields
answers to the wrong questions ("I know you heard
what I said, but did you understand what I meant?").
Some of the questions managers should ask
technical support staff, and vice versa, to ensure that
the solution being developed is appropriate to the
actual problems are listed in Table 8-6 through 8-8.
Table 8-6 consists of "screening level" questions.
Table 8-7 addresses the need for correct
conceptualizations, and Table 8-8 is comprised of
sociopolitical concerns.
On another level of communication, managers should
appreciate how difficult it will be to explain the results
of complicated models to non-technical audiences
such as in public meetings and courts of law. Many
scientists find it a trying exercise to discuss the
details of their labors without the convenience of the
jargon of their discipline. Some of the more useful
means of overcoming this limitation involve the
production of highly simplified audio-visual aids, but
this necessarily involves a great deal of work. The
efforts which will be required to sell purportedly self-
explanatory graphs from computer simulations may
rival the efforts spent on producing the simulations
initially.
Table 8-6 Screening-Level Questions for Mathematical
Modeling Efforts
General Problem Definition
What are the key issues; quantity, quality, or both?
What are the controlling geologic, hydrologic, chemical,
and biological features?
Are there reliable data (proper field scale, quality controlled,
etc.) for preliminary assessments?
Do we have the model(s) needed for appropriate
simulations?
Initial Responses Needed
What is the time-frame for action (imminent or long-term)?
What actions, if taken now, can significantly delay the
projected impacts?
To what degree can mathematical simulations yield
meaningful results for the action alternatives, given
available data?
What other techniques or information (generic models, past
experience, etc.) would be useful for initial estimates?
Strategies for Further Study
Are the critical data gaps identified; if not, how well can
simulations determine the specific data needs?
What are the trade-offs between additional data and
increased certainty of the simulations?
How much additional manpower and resources are
necessary for further modeling efforts?
How long will it take to produce useful simulations,
including quality control and error-estimation efforts?
Table 8-7 Conceptualization Questions for Mathematical
Modeling Efforts
Assumptions and Limitations
What are the assumptions made, and do they cast doubt on the
model's projections for this problem?
What are the model's limitations regarding the natural processes
controlling the problem; can the full spectrum of probable
conditions be addressed?
How far in space and time can the results of the model
simulations be extrapolated?
Where are the weak spots in the application, and can these be
further minimized or eliminated?
Input Parameters and Boundary Conditions
How reliable are the estimates of the input parameters; are they
quantified within accepted statistical bounds?
What are the boundary conditions, and why are they appropriate
to this problem?
Have the initial conditions with which the model is calibrated been
checked for accuracy and internal consistency?
Are the spatial grid design(s) and time-steps of the model
optimized for this problem?
Quality Control and Error Estimation
Have these models been mathematically validated against other
solutions to this kind of problem?
Has anyone field verified these models before, by direct
applications or simulation of controlled experiments?
How do these models compare with others in terms of
computational efficiency, and ease of use or modification?
What special measures are being taken to estimate the overall
errors of the simulations?
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Table 8-8 Sociopolitical Questions for Mathematical
Modeling Efforts
Demographic Considerations
Is there a larger population endangered by the problem than we
are able to provide sufficient responses to?
Is it possible to present the model's results in both nontechnical
and technial formats, to reach all audiences?
What role can modeling play in public information efforts?
How prepared are we to respond to criticism of the model(s)?
Political Constraints
Are there nontechnical barriers to using this model, such as
"tainted by association" with a controversy elsewhere?
Do we have the cooperation of all involved parties in obtaining the
necessary data and implementing the solution?
Are similar technical efforts for this problem being undertaken by
friend or foe?
Can the results of the model simulations be turned against us; are
the results ambiguous or equivocal?
Legal Concerns
Will the present schedule allow all regulatory requirements to be
met in a timely manner?
If we are dependent on others for key inputs to the model(s), how
do we recoup losses stemming from their nonperformance?
What liabilities are incurred for projections which later turn out to
be misinterpretations originating in the model?
Do any of the issues relying on the application of the model(s)
require the advice of attorneys?
8.5 References
Aller, L, T. Bennett, J.H. Lehr, and R.J. Petty. 1985.
DRASTIC: A Standardized System for Evaluating
Ground Water Pollution Potential Using
Hydrogeologic Settings. EPA-600/2-85-018, U.S.
Environmental Protection Agency, Robert S. Kerr
Environmental Research Laboratory, Ada, OK.
Andersen, P.P., C.R. Faust and J.W. Mercer. 1984.
Analysis of Conceptual Designs for Remedial
Measures at Lipari Landfill. Ground Water 22(2):
176-190.
Asbeck, E., and Y.Y. Haimes. 1984. The Partitioned
Multiobjective Risk Method. Large Scale Systems
13(38).
Bachmat, Y., B. Andrews, D. Holtz, and S. Sebastian.
1978. Utilization of Numerical Groundwater Models
for Water Resource Management. EPA-600/ 8-
78-012, U.S. Environmental Protection Agency,
Robert S. Kerr Environmental Research Laboratory,
Ada, OK.
Boonstra, J., and N.A. de Ridder. 1981. Numerical
Modeling of Groundwater Basins. ILRI Publication No.
29, International Institute for Land Reclamation and
Improvement, Wageningen, The Netherlands.
Brown, J. 1986. 1986 Environmental Software
Review. Pollution Engineering 18(1):18-28.
Daubert, J.T., and R.A. Young. 1982. Ground-Water
Development in Western River Basins: Large
Economic Gains with Unseen Costs. Ground Water
20(1):80-86.
Domenico, P.A. 1972. Concepts and Models in
Groundwater Hydrology. McGrawHill, New York, NY
Faust, C.R., LR. Silka, and J.W. Mercer. 1981.
Computer Modeling and Ground-Water Protection.
Ground Water 19(4):362-365.
Freeze, R.A., and J.A. Cherry. 1979. Groundwater.
Prentice Hall, Englewood Cliffs, NJ.
Graves, B. 1986. Ground Water Software -
Trimming the Confusion. Ground Water Monitoring
Review 6(1):44-53.
Haimes, Y.Y., ed. 1981. Risk/Benefit Analysis in
Water Resources Planning and Management.
Plenum Publishers, New York, NY.
Haimes, Y.Y., and W.A. Hall. 1974. Multiobjectives in
Water Resources Systems Analysis: The Surrogate
Worth Trade-off Method. Water Resources
Research 10:615-624.
Heath, R.C. 1982. Classification of Ground-Water
Systems of the United States. Ground Water
20(4):393-401.
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Holcolm Research Institute. 1976. Environmental
Modeling and Decision Making. Praeger Publishers,
New York, NY.
Huyakorn, P.S., A.G. Kretschek, R.W. Broome, J.W.
Mercer, and B.H. Lester. 1984. Testing and Validation
of Models for Simulating Solute Transport in Ground
Water: Development, Evaluation, and Comparison of
Benchmark Techniques. IGWMC Report No. GWMI
84-13, International Ground Water Modeling Center,
Holcolm Research Institute, Butler University.
International Ground Water Modeling Center. 1986.
Price List of Publications and Services Available from
IGWMC (January 1986). International Ground Water
Modeling Center, Holcolm Research Institute, Butler
University.
Javendel, I., C. Doughty, and C.F. Tsang. 1984.
Groundwater Transport: Handbook of Mathematical
Models. AGU Water Resources Monograph No. 10.,
American Geophysical Union, Washington, DC.
Kazmann, R.G. 1972. Modern Hydrology. 2nd ed.
Harper & Row Publishers, New York, NY.
Khan, I.A. 1986a. Inverse Problem in Ground Water:
Model Development. Ground Water 24(1):32-38.
Khan, I.A. 1986b. Inverse Problem in Ground Water:
Model Application. Ground Water 24(1):39-48.
Krabbenhoft, D.P., and M.P. Anderson. 1986. Use of
a Numerical Ground-Water Flow Model for
Hypothesis Testing. Ground Water 24(1):49-55.
Lehr, J.H. 1986. The Myth of TVA. Ground Water
24(1):2-3.
Lohman, S.W. 1972. Ground-Water Hydraulics.
U.S. Geological Survey Professional Paper 708, U.S.
Government Printing Office, Washington, D C.
Mercer, J.W., and C.R. Faust. 1981. Ground-Water
Modeling. National Water Well Association,
Worthington, OH.
Miller, S. 1980. Cost-Benefit Analyses.
Environmental Science and Technology
14(12):1415-1417.
Molz, F.J., O. Guven, and J.G. Melville. 1983. An
Examination of Scale Dependent Dispersion
Coefficients. Ground Water 21(6):715-725.
Moses, C.O., and J.S. Herman. 1986. Computer
Notes - WATIN - A Computer Program for
Generating Input Files for WATEQF. Ground Water
24(1):83-89.
Puri, S. 1984. Aquifer Studies Using Flow
Simulations. Ground Water 22(5):538-543.
Remson, I., G.M. Hornberger, and F.J. Molz. 1971.
Numerical Methods in Subsurface Hydrology. John
Wiley and Sons, New York, NY.
Ross, B., J.W. Mercer, S.D. Thomas, and B.H.
Lester. 1982. Benchmark Problems for Repository
Siting Models. U.S. NRC Publication No. NUREG/
CR-3097, U.S. Nuclear Regulatory Commission,
Washington, DC.
Shelton, M.L. 1982. Ground-Water Management in
Basalts. Ground Water 20(1):86-93.
Srinivasan, P. 1984. PIG - A Graphic Interactive
Preprocessor for Ground-Water Models. IGWMC
Report No. GWMI 84-15, International Ground Water
Modeling Center, Holcolm Research Institute, Butler
University.
Strecker, E.W., and W. Chu. 1986. Parameter
Identification of a Ground-Water Contaminant
Transport Model. Ground Water 24(1):56-62.
Todd, O.K. 1980. Groundwater Hydrology. 2nd ed.
John Wiley and Sons. New York.
U.S. Congress. 1982. Use of Models for Water
Resources Management, Planning, and Policy. Office
of Technology Assessment U.S. Government Printing
Office, Washington, DC.
U.S. Environmental Protection Agency. 1984.
Ground-Water Protection Strategy. Office of
Ground-Water Protection, Washington, DC.
van der Heijde, P.K.M. 1985. The Role of Modeling in
Development of Ground-Water Protection Policies.
Ground Water Modeling Newsletter 4(2).
van der Heijde, P.K.M., Y. Bachmat, J. Bredehoeft, B.
Andrews, D. Holtz, and S. Sebastian. 1985.
Groundwater Management: The Use of Numerical
Models. 2nd ed. AGU Water Resources Monograph
No. 5, American Geophysical Union, Washington, DC.
van der Heijde, P.K.M. 1984a. Availability and
Applicability of Numerical Models for Ground Water
Resources Management. IGWMC Report No. GWMI
84-14, International Ground Water Modeling Center,
Holcolm Research Institute, Butler University.
van der Heijde, P.K.M. 1984b. Utilization of Models as
Analytic Tools for Groundwater Management. IGWMC
Report No. GWMI 84-19, International Ground Water
Modeling Center, Holcolm Research Institute, Butler
University.
van der Heijde, P.K.M., and P. Srinivasan. 1983.
Aspects of the Use of Graphic Techniques in Ground
Water Modeling. IGWMC Report No. GWMI 83-11,
International Ground Water Modeling Center, Holcolm
Research Institute, Butler University.
Wang, H.F., and M.P. Anderson. 1982. Introduction to
Groundwater Modeling: Finite Difference and Finite
Element Methods. W.H. Freeman and Company, San
Francisco, CA.
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Warren, J., H.P. Mapp, D.E. Ray, D.D. Kletke, and C.
Wang. 1982. Economics of Declining Water Supplies
of the Ogallala Aquifer. Ground Water 20(1):73-80.
White, J.A., M.H. Agee, and K.E. Case. 1984.
Principles of Engineering Economic Analysis. 2nd ed.
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CHAPTER 9
BASIC GEOLOGY
Geology, the study of the earth, includes the
investigation of earth materials, the processes that act
on these materials, the products that are formed, the
history of the earth, and the origin and development
of life forms. There are several subfields of geology.
Physical geology deals with all aspects of the earth
and includes most earth science specialties. Historical
geology is the study of the origin of the earth,
continents and ocean basins, and life forms, while
economic geology is an applied approach involved in
the search and exploitation of mineral resources,
such as metallic ores, fuels, and water. Structural
geology deals with the various structures of the earth
and the forces that produce them. Geophysics is the
examination of the physical properties of the earth
and includes the study of earthquakes and methods
to evaluate the subsurface.
From the perspective of ground water, all of the
subfields of geology are used, some more than
others. Probably the most difficult concept to
comprehend by individuals with little or no geological
training is the complexity of the subsurface, which is
hidden from view and, at least presently, cannot be
adequately sampled. In geologic or hydrogeologic
studies, it is best to always keep in mind a
fundamental principle of geology. The present is the
key to the past. That is, the processes occurring
today are the same processes that occurred
throughout the geologic past -- only the magnitude
changes from one time to the next.
Consider, for example, the channel and flood plain of
a modern day river or stream. The watercourse
constantly meanders from one side of the flood plain
to the other, eroding the banks and carrying the
sediments farther downstream. The channel changes
in size and position, giving rise to deposits of differing
grain size and, perhaps, composition. The changes
may be abrupt or gradual, both vertically and
horizontally, as is evident from an examination of the
walls of a gravel pit or the bluffs along a river.
Because of the dynamic nature of streams and
deltas, one will find a geologic situation that is
perplexing not only to the individual involved in a
ground-water investigation, but to the geologist as
well. Each change in grain size will cause differences
in permeability and ground-water velocity, while
changes in mineral composition can lead to variances
in water quality. At the other end of the depositional
spectrum are deposits collected in lakes, seas, and
the oceans, which are likely to be much more
widespread and uniform in thickness, grain size, and
composition.
As one walks from the sandy beach of a lake into the
water, the sediments become finer and more widely
distributed as the action of waves and currents sort
the material brought into the lake by streams. Farther
from shore, the bottom of the lake may consist of
mud, which is a mixture of silt, clay, and organic
matter. In some situations the earthy mud grades
laterally into a lime ooze or mud. In geologic time,
these sediments become lithified or changed into rock
~ the sand to sandstone, the mud to shale, and
the limey mud to limestone. It is important to note,
however, that the sand, mud, and lime were all
deposited at the same time, although with lithification
each sediment type produced a different sedimentary
rock.
9.1 Geologic Maps and Cross-Sections
Geologists use a number of techniques to graphically
represent surface and subsurface conditions. These
include surficial geologic maps, columnar sections,
cross-sections of the subsurface, maps that show
the configuration of the surface of a geologic unit,
such as the bedrock beneath glacial deposits, maps
that indicate the thickness or grain size of a particular
unit, a variety of contour maps, and a whole host of
others.
A surficial geologic map depicts the geographic extent
of formations and their structure. Columnar sections
describe the vertical distribution of rock units, their
lithology, and thickness. Geologic cross-sections
attempt to illustrate the subsurface distribution of rock
units between points of control, such as outcrops or
well bores. An isopach map shows the geographic
range in thickness of a unit; these maps are based
largely or entirely on well logs.
Whatever the graphic techniques, it must be
remembered that these maps represent only best
guesses and may be based on scanty data. In reality,
they are interpretations, presumably based on
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scientific thought, a knowledge of depositional
characteristics of rock units, and a data base that
provides some control. They are not exact because
the features they attempt to show are complex, nearly
always hidden from view, and difficult to sample.
All things considered, graphical representations are
exceedingly useful, if not essential, to subsurface
studies. On the other hand, a particular drawing that
is prepared for one purpose may not be adequate for
another even though the same units are involved.
This is largely due to scale and generalizations.
A combination topographic and geologic map of a
glaciated area is shown in Figure 9-1. The upland
area is mantled by glacial till (Qgm) and the surficial
material covering the relatively flat flood plain has
been mapped as alluvium (Qal). Beneath the alluvial
cover are other deposits of glacial origin that consist
of glacial till, outwash, and local lake deposits. A
water well drillers log of a boring at point A states,
"This well is just like all of the others in the valley,"
and that the upper 70 feet of the valley fill consists of
a "mixture of clay, sand, silt, and boulders." This is
underlain by 30 feet of "water sand," which is the
aquifer. The aquifer overlies "slate, jingle rock, and
coal." The terminology may be quaint, but it is
nonetheless a vocabulary that must be interpreted.
Examination of the local geology, as evidenced by
strata that crop out along the hillsides, indicate that
the bedrock or older material that underlies the glacial
drift consists of shale, sandstone cemented by
calcite, and lignite, which is an immature coal. These
are the geologic terms, at least in this area, for "slate,
jingle rock, and coal," respectively.
For generalized purposes, it is possible to use the
driller's log to construct a cross section across or
along the stream valley (Figure 9-2). In this case,
one would assume for the sake of simplicity the
existence of an aquifer that is rather uniform in
composition and thickness. A second generation
cross section, shown in Figure 9-3, is based on
several bore hole logs described by a geologist who
collected samples as the holes were being drilled.
Notice in this figure that the subsurface appears to be
much more complex, consisting of several isolated
permeable units that are incorporated within the fine
grained glacial deposits that fill the valley. In addition,
the aquifer does not consist of a uniform thickness of
sand, but rather a unit that ranges from 30 to 105 feet
in thickness and from sand to a mixture of sand and
gravel. The water-bearing characteristics of these
units are all different. This cross section too is quite
generalized, which becomes evident as one examines
an actual log of one of the holes (Table 9-1).
Figure 9-1
Generalized geologic map of a glaciated area
along the Souris River Valley in central North
Dakota.
Qgm
x^ Scale (miles)
Qal = alluvium
Qgm = ground moraine
Qkt = terrace deposits
Figure 9-2
Generalized geologic cross section of the
Souris River Valley based on driller's log.
Water Well
186
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Table 9-1 Geologist's Log of a Test Hole, Souris River Table 9-2
Valley, North Dakota
Sample Description
and Drilling Condition Depth (ft)
Generalized Geologic Logs of Five Test
Holes, Souris River Valley, North Dakota
Material
Topsoil, silty clay, black 0-1
Clay, silty, yellow brown, poorly consolidated 1-5
Clay, silty, yellow grey, soft, moderately compacted 5-10
Clay, silty, as above, silty layers, soft 10-15
Silty, clayey, gray, soft, uniform drilling 15-20
Clay, silty, some fine to medium sand, gray 20-30
Clay, gray to black, soft, very tight 30-40
Clay, as above, gravelly near top 40-50
Clay, as above, no gravel 50-60
Clay, as above, very silty in spots, gray 60-70
Clay and silt, very easy drilling 70-80
Clay, as above to gravel, fine to coarse,
sandy, thin clay layers, taking lots of water 80-90
Gravel, as above, some clay near top, very
rough drilling, mixed three bags of mud, lots
of lignite chips 90-100
Gravel, as above, cobbles and boulders 100-120
Gravel, as above, to sand, fine to coarse,
lots of lignite, much easier drilling 120-130
Clay, gravelly and rocky, rough drilling, poor
sample return 130-140
Sandy clay, gravelly and rocky, rough drilling,
poor sample return (till) 140-150
Sandy clay, as above, poor sample return 150-160
Clay, sandy, gray, soft, plastic, noncalcareous 160-170
Clay, sandy, as above, tight, uniform drilling 170-180
Clay, as above, much less sand, gray, soft,
tight, plastic 180-190
Clay, as above, no sand, good sample return 190-200
Clay, as above 200-210
In addition to showing more accurately the
composition of the subsurface, logs can also provide
some interesting clues concerning the relative
permeabilities of the water-bearing units. Referring
to Table 9-2, the depth interval ranging from 62 to
92 feet, a generalized log of well 1 contains the
remark "losing water" and in well 5, at depths of 80
to 120 feet, is the notation, "3 bags of bentonite." In
the first case this means that the material being
penetrated by the drill bit from 62 to 92 feet was more
permeable than the annulus of the cutting-filled bore
hole. The water, pumped down the hole through the
drill pipe to remove the cuttings, found it easier to
move out into the formation than to flow back up the
hole. The remark is a good indication of a
permeability that is higher than that present in those
sections where water was not being lost.
In the case of well 5, the material was so permeable
that much of the drilling fluid was moving into the
formation and there was no return of the cuttings. To
regain circulation, bentonite, or to use the field term,
mud, was added to the drilling fluid to seal the
permeable zone. Even though the geologist described
Depth (ft)
Test Hole 1
Fill 0-3
Silt, olive-gray 3-14
Sand, fine-medium 14-21
Silt, sandy, gray 21-25
Clay, gray 25-29
Sand, fine-coarse 29-47
Clay, gray 47-62
Gravel, fine to coarse, losing water 62-92
Silt, sandy, gray 92-100
Observation well depth 80 feet
Test Hole 2
Fill 0-2
Clay, silty and sandy, gray 2-17
Clay, very sandy, gray 17-19
Sand, fine-medium 19-60
Sand, fine-coarse with gravel 60-80
Gravel, coarse, 2 bags bentonite and bran 80-100
Observation well depth 88 feet
Test Hole 3
Silt, yellow 0-5
Clay, silty, black 5-15
Sand, fine to coarse 15-29
Clay, silty, gray 29-65
Sand, medium-coarse, some gravel 65-69.
Gravel, sandy, taking water 69-88
Sand, fine to medium, abundant chips of lignite 88-170
Observation well depth 84 feet
Test Hole 4
Fill 0-5
Silt, brown 5-12
Sand, fine-medium 12-28
Clay, silty and sandy, gray 28-37
Sand, fine 37-49
Clay, dark gray 49-55
Sand, fine 55-61
Clay, sandy, gray 61-66
Sand, fine-coarse, some gravel 66-103
Silt, gray 103-120
Observation well depth 96 feet
Test Hole 5
Clay, silty, brown 0-10
Silt, clayey, gray 10-80
Gravel, fine-coarse, sandy, taking lots of water
3 bags bentonite 80-120
Sand, fine to coarse, gravelly 120-130
Clay, gravelly and rocky (till) 130-150
Sand, fine, Fort Union Group 150-180
Observation well depth 100 feet
187
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Figure 9-3 Geologic cross section of the Souris River Valley based on detailed logs of test holes.
Test Test Test
Hole Hole Hole
Test
Hole
— -- -Clay— S
ilU^l: S !•,-.•:-.;:^
• . °-. :• • » o- • *:—•. • •» ••••'*:
•'•<>."'•••• »•« • . • ." . .o • .• - •-'.*.•" f-
JJIl^^^iiilS ^--r i'J-v°i •''.»•/>'••: Vv^-y--; ,•'••?•• ^•••?:
IF?^"*-^^ sit° "Ji^~^-;^''^:^4y 2?!'°.'-;0«j;.''O
•'^•':''V:"^'^^^\''^'-^'°-^'''':V'-^.l^'^I''t>'^^^
Figure 9-4 Schematic of general features of the Columbia Plateau region (from Heath, 1984).
Older Mountains
River Canyon
Present soil zone
Interflow zone
Silt and clay
Cooling fractures
188
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the aquifer materials from both zones similarly, the
section in well 5 is more permeable than the one in
well 1, which in turn is more permeable than the other
coarse-grained units penetrated where there was no
fluid loss.
The three most important points to be remembered
here are, first, that graphical representations of the
surface or subsurface geology are merely guesses of
what might actually occur, and even these depend to
some extent on the original intended usage. Second,
the subsurface is far more complex than is usually
anticipated, particularly in regard to unconsolidated
deposits. Finally, evaluating the original data such as
well logs might lead to a better evaluation of the
subsurface, an evaluation that far surpasses the use
of generalized lithologic logs alone.
9.2 Ground Water in Igneous and
Metamorphic Rocks
Nearly all of the porosity and permeability of igneous
and metamorphic rocks is the result of secondary
openings such as fractures, faults, and the dissolution
of certain minerals. A few notable exceptions include
large lava tunnels present in some flows, interflow or
coarse sedimentary layers between individual lava
flows, and deposits of selected pyroclastic materials
(Figure 9-4).
Because the openings in igneous and metamorphic
rocks are quite small volumetrically, rocks of this type
are poor suppliers of ground water. The supplies that
are available commonly drain rapidly after a period of
recharge by infiltration of precipitation. In addition they
are subject to contamination from the surface where
these rocks crop out.
The width, spacing, and depth of fractures range
widely, as do their origins, the surface to 0.003
inches at a depth of 200 feet, while spacing increased
from 5 to 10 feet near the surface to 15 to 35 feet at
depth in the Front Range of the Rocky Mountains. He
also reported that porosity decreased from below 300
feet or so, but there are many recorded exceptions.
Exfoliation fractures in the crystalline rocks of the
Piedmont near Atlanta, Georgia range from 1 to 8
inches in width (Cressler and others, 1983).
The difficulty of evaluating water and contaminant
movement in fractured rocks is that the actual
direction of movement may not be in the direction of
decreasing head, but rather in some different though
related direction. The problem is further compounded
by the difficulty in locating the fractures. Because of
these characteristics, evaluation of water availability,
direction of movement, and velocity is exceedingly
difficult. As a general rule, at least in the eastern part
of the United States, LeGrand pointed out that well
yields, and therefore fractures, permeability, and
porosity, are greater in valleys and broad ravines than
on flat uplands, which in turn are higher than on hill
slopes and hill crests (Figure 9-5). The reason this
occurs in parts of North Carolina is because stream
valleys have formed along fracture zones.
Unless some special circumstance exists, such as
where rocks crop out at the surface, water obtained
from igneous and metamorphic rocks is nearly always
of excellent chemical quality. Dissolved solids are
commonly less than 100 mg/l. Water from
metamorphosed carbonate rocks may have moderate
to high concentrations of hardness.
9.3 Ground Water in Sedimentary Rocks
Usable supplies of ground water can be obtained
from all types of sedimentary rocks, but the fine
grained strata such as shale and siltstone may only
provide a few gallons per day and even this can be
highly mineralized. Even though fine grained rocks
may have relatively high porosities, the primary
permeability is very low. On the other hand, shale is
likely to contain a great number of joints that are both
closely spaced and extend to considerable depths.
Therefore, rather than being impermeable, as many
individuals imply, they can be quite transmissive. This
is of considerable importance in waste disposal
schemes because insufficient attention might be paid
during engineering design to the potential for flow
through fractures. In addition, the leachate that is
formed as water infiltrates through waste might be
small in quantity but highly mineralized. Because of
the low bulk permeability, it would be difficult to pump
out the contaminated water or even to properly locate
monitoring wells.
From another perspective, fine grained sedimentary
rocks, owing to their high porosity, can store huge
quantities of water. Some of this water can be
released to adjacent aquifers when a head difference
is developed due to pumping. No doubt fine grained
confining units provide, on a regional scale, a great
deal of water to aquifer systems. The porosity,
however, decreases with depth because of
compaction brought about by the weight of overlying
sediments.
The porosity of sandstones range from less than 1
percent to a maximum of about 30 percent. This is a
function of sorting, grain shape, and cementation.
Cementation can be variable both in space and time
and on outcrops can differ greatly from that in the
subsurface.
Fractures also play an important role in the movement
of fluids through sandstones and transmissivities may
be as much as two orders of magnitude greater in a
fractured rock than in an unfractured part of the same
geologic formation.
Sandstone units that were deposited in a marine or
near-marine environment can be very widespread,
covering tens of thousands of square miles, such as
the St. Peter sandstone of Cambrian age. Those
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Figure 9-5 Schematic Of general features of the Piedmont and Blue Ridge region (from Heath, 1984).
Bedrock Outcrops
Figure 9-6 Schematic of general features of the Gulf Coastal Plain (from Heath, 1984).
Parallel Outcrop Belt
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representing ancient alluvial channel fills, deltas, and
related environments of deposition are more likely to
be discontinuous and erratic in thickness. Individual
units are exceedingly difficult to trace in the
subsurface. Regional ground-water flow and storage
may be strongly influenced by the geologic structure
(Figures 9-6 and 9-7).
Carbonate rocks are formed in many different
environments and the original porosity and
permeability are modified rapidly after burial. Some
special carbonate rocks, such as coquina and some
breccias, may remain very porous and permeable, but
these are the exception.
It is the presence of fractures and other secondary
openings that develop high yielding carbonate
aquifers. One important aspect is the change from
calcite to dolomite (CaMgCOa), which results in a
volumetric reduction of 13 percent and the creation of
considerable pore space. Of particular importance
and also concern in many of the carbonate regions of
the world is the dissolution of carbonates along
fractures and bedding planes by circulating ground
water. This is the manner in which caves and
sinkholes are formed. As dissolution progresses
upward in a cave, the overlying rocks may collapse to
form a sinkhole that contains water if the cavity
extends below the water table. Regions in which there
has been extensive dissolution of carbonates leading
to the formation of caves, underground rivers, and
sinkholes, are called karst. Notable examples include
parts of Missouri, Indiana, and Kentucky (Figure 9-
8).
Karst areas are particularly troublesome even though
they can provide large quantities of water to wells and
springs because they are easily contaminated, it is
often difficult to trace the contaminant, the water can
flow very rapidly, and there is no filtering action to
degrade the waste. Not uncommonly a well owner
may be unaware that he is consuming unsafe water.
9.4 Ground Water in Unconsolidated
Sediments
Unconsolidated sediments accumulate in many
different environments, all of which leave their mark
on the characteristics of the deposit. Some are thick
and areally extensive, as the alluvial fill in the Basin
and Range Province; others are exceedingly long and
narrow, such as the alluvial deposits along streams
and rivers; and others may cover only a few hundred
square feet, for example, some glacial forms. In
addition to serving as major aquifers, Unconsolidated
sediments are also important as sources of raw
materials for construction.
Although closely related to sorting, the porosities of
Unconsolidated materials range from less than 1 to
more than 90 percent, the latter representing
uncompacted mud. Permeabilities also range widely.
Cementing of some type and degree is probably
universal, but not obvious, with silt and clay being the
predominant form.
Most unconsolidated sediments owe their
emplacement to running water and consequently,
some sorting is expected. On the other hand, water
as an agent of transportation will vary in both volume
and velocity, which is climate dependent, and this will
leave an imprint on the sediments. It is to be
expected that stream related material, which most
unconsolidated material is, will be variable in extent,
thickness, and grain size (Figure 9-9). Other than
this, one can draw no general guidelines; therefore, it
is essential to develop some knowledge of the
resulting stratigraphy that is characteristic of the most
common environments of deposition. The water-
bearing properties of glacial drift, of course, are
exceedingly variable, but stratified drift is more
uniform and better sorted than glacial till (Figure 9-
10).
9.5 Relationship Between Geology,
Climate, and Ground-Water Quality
The availability of ground-water supplies and their
chemical quality are closely related to precipitation.
As a general rule, the least mineralized water, both in
streams and underground, occurs in areas of the
greatest amount of rainfall. Inland, precipitation
decreases, water supplies diminish, and quality
deteriorates. Water-bearing rocks exert a strong
influence on ground-water quality and thus, the
solubility of the rocks may override the role of
precipitation.
Where precipitation exceeds 40 inches per year,
shallow ground water usually contains less than 500
mg/l and commonly less than 250 mg/l of dissolved
solids. Where precipitation ranges between 20 and 40
inches, dissolved solids may range between 400 and
1,000 mg/l, and in drier regions dissolved solids
commonly exceed 1,000 mg/l.
The dissolved solids concentration of ground water
increases toward the interior of the continent. The
increase is closely related to precipitation and the
solubility of the aquifer framework. The least
mineralized ground water is found in a broad belt that
extends southward from the New England States
along the Atlantic Coast to Florida, and then
continues to parallel much of the Gulf Coast.
Similarly, along the Pacific Coast from Washington to
central California the mineral content is also very low.
Throughout this belt, dissolved solids concentrations
are generally less than 250 mg/l and commonly less
than 100 mg/l (Figure 9-11).
The Appalachian region consists of a sequence of
strata that range from nearly flat-lying to complexly
folded and faulted. Likewise, ground-water quality in
this region is also highly variable, being generally
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Figure 9-7 Schematic of general features of the Colorado Plateau and Wyoming Basin region (from Heath, 1984).
Extinct Volcanoes
Ridges
Dome
Canyon
Cliff
Fault Scarp
Fault
Fresh water
Salty water
Sandstone
i ' i I Limestone
•_-_-3 Shale
Metamorphic
rocks
Figure 9-8 Schematic of general features of the Nonglaciated Central region (from Heath, 1984).
Regolith
Fresh water
Salty water
192
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Figure 9-9 Schematic of general features of the High Plains region (from Heath, 1984).
Platte River
Figure 9-10 Schematic of general features of the Glaciated Central region (from Heath, 1984).
Loess
Fresh water
Salty water
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Figure 9-11 Dissolved solids concentrations in ground water used for drinking in the United States (from Pettyjohn and
others, 1979).
Dissolved solids concentrations, mg/l
| |<250
JM 250-500
500-1000
> 1000
0 100 200
Scale (miles)
harder and containing more dissolved minerals than
does water along the coastal belt. Much of the
difference in quality, however, is related to the
abundance of carbonate aquifers which provide
waters rich in calcium and magnesium.
Westward from the Appalachian Mountains to about
the position of the 20 inch precipitation line (eastern
North Dakota to Texas), dissolved solids in ground
water progressively increase. They are generally less
than 1000 mg/l and are most commonly in the 250 to
750 mg/l range. The water is moderately to very hard,
and in some areas concentrations of sulfate and
chloride are excessive.
From the 20 inch precipitation line westward to the
northern Rocky Mountains, dissolved solids are in the
500 to 1500 mg/l range. Much of the water from
glacial drift and bedrock formations is very hard and
contains significant concentrations of calcium sulfate.
Other bedrock formations may contain soft sodium
bicarbonate, sodium sulfate, or sodium chloride water.
Throughout much of the Rocky Mountains, ground-
water quality is variable, although Ihe dissolved solids
concentrations commonly range between 250 and
750 mg/l. Stretching southward from Washington to
southern California, Arizona, and New Mexico is a
vast desert region. Here the difference in quality is
wide and dissolved solids generally exceed 750 mg/l.
In the central parts of some desert basins the ground
water is highly mineralized, but along the mountain
flanks the mineral content may be quite low.
Extremely hard water is found over much of the
interior lowlands, Great Plains, Colorado Plateau, and
Great Basin. Isolated areas of high hardness are
present in northwestern New York, eastern North
Carolina, the southern tip of Florida, northern Ohio,
and parts of southern California. In general, the
hardness is of the carbonate type.
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On a regional level, chloride does not appear to be a
significant problem, although it is troublesome locally
due largely to industrial activities, the intrusion of
seawater caused by overpumping coastal aquifers, or
interaquifer leakage related to pressure declines
brought about by withdrawals.
In many locations, sulfate levels exceed the Federal
recommended limit of 250 mg/l; regionally, sulfate
may be a problem only in the Great Plains, eastern
Colorado Plateau, Ohio, and Indiana. Iron problems
are ubiquitous; concentrations exceeding only 0.3
mg/l will cause staining of clothing and fixtures.
Fluoride is abnormally high in several areas,
particularly parts of western Texas, Iowa, Illinois,
Indiana, Ohio, New Mexico, Wyoming, Utah, Nevada,
Kansas, New Hampshire, Arizona, Colorado, North
and South Dakota, and Louisiana.
A water-quality problem of growing concern,
particularly in irrigated regions, is nitrate, which is
derived from fertilizers, sewage, and through natural
causes. When consumed by infants less than six
months old for a period of time, high nitrate
concentrations can cause a disease known as "blue
babies." This occurs because the child's blood
cannot carry sufficient oxygen; the disease is easily
overcome by using low nitrate water for formula
preparation. Despite the fact that nitrate
concentrations in ground water appear to have been
increasing in many areas during the last 30 years or
so, the concern may be more imagined than real
because there have been no reported incidences of
"blue babies" for more than 20 years, at least in the
states that comprise the Great Plains.
In summary, the study of geology is complex in detail,
but the principles outlined above should be sufficient
for a general understanding of the topic, particularly
as it relates to ground water. If interested in a more
definitive treatment, the reader should refer to the
following sections and the references at the end of
the chapter.
9.6 Minerals
The earth, some 7,926 miles in diameter at the
equator, consists of a core, mantle, and crust, which
have been defined by the analysis of seismic or
earthquake waves. Only a thin layer of the crust has
been examined by humans. It consists of a variety of
rocks, each of which is made up of one or more
minerals.
Most minerals contain two or more elements, but of
all the elements known, eight account for nearly 98
percent of the rocks and minerals:
Oxygen 46%
Silicon 27.72%
Aluminum 8.13%
Iron 5%
Calcium 3.63%
Sodium 2.83%
Potassium 2.59%
Magnesium 2.09%
Without detailed study, it is usually difficult to
distinguish one mineral from another, except for a few
common varieties such as quartz, pyrite, mica, and
some gemstones. On the other hand, it is important
to have at least a general understanding of
mineralogy because it is the mineral make-up of
rocks that, to a large extent, controls the type of
water that a rock will contain under natural conditions
and the way it will react to contaminants or naturally
occurring substances.
The most common rock-forming minerals are
relatively few and deserve at least a mention. They
can be divided into three broad groups: (1) the
carbonates, sulfates, and oxides; (2) the rock-
forming silicate minerals; and (3) the common ore
minerals.
9.6.7 Carbonates, Sulfates, and Oxides
Calcite, a calcium carbonate (CaCOa), is the major
mineral in limestone. The most common mineral is
quartz. It is silicon dioxide (SiO2), hard, and resistant
to both chemical and mechanical weathering. In
sedimentary rocks it generally occurs as sand-size
grains (sandstone) or even finer, such as silt or clay
size, and it may also appear as a cement. Because of
the low solubility of silicon, silica generally appears in
concentrations less than 25 mg/l in water. Limonite is
actually a group name for the hydrated ferric oxide
minerals (Fe203*H20), which occur so commonly in
many types of rocks. Limonite is generally rusty or
blackish with a dull, earthy luster and a yellow-brown
streak. It is a common weathering product of other
iron minerals. Because limonite and other iron-
bearing minerals are nearly universal, dissolved iron is
a very common constituent in water, causing staining
of clothing and plumbing fixtures. Gypsum, a hydrated
calcium sulfate (CaSO4»2H20), occurs as a
sedimentary evaporite deposit and as crystals in shale
and some clay deposits. Quite soluble, it is the major
source of sulfate in ground water.
9.6.2 Rock-Forming Silicates
The most common rock-forming silicate minerals
include the feldspars, micas, pyroxenes, amphiboles,
and olivine. Except in certain igneous and
metamorphic rocks these minerals are quite small and
commonly require a microscope for identification. The
feldspars are alumino-silicates of potassium or
sodium and calcium. Most of the minerals in this
group are white, gray, or pink. Upon weathering they
turn to clay and release the remaining chemical
elements to water. The micas muscovite and biotite
are platy alumino-silicate minerals that are common
and easily recognized in igneous, metamorphic, and
sedimentary rocks. The pyroxenes, a group of
silicates of calcium, magnesium, and iron, as well as
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the amphiboles, which are complex hydrated silicates
of calcium, magnesium, iron, and aluminum, are
common in most igneous and metamorphic rocks.
They appear as small, dark crystals of accessory
minerals. Olivine, a magnesium-iron silicate, is
generally green or yellow and is common in certain
igneous and metamorphic rocks. None of the rock-
forming silicate minerals have a major impact on
water quality in most situations.
9.6.3 Ores
The three most common ore minerals are galena,
sphalerite, and pyrite. Galena, a lead sulfide (PbS), is
heavy, brittle, and breaks into cubes. Sphalerite is a
zinc sulfide (ZnS) mineral that is brownish, yellowish,
or black. It ordinarily occurs with galena and is a
major ore of zinc. The iron sulfide pyrite (FeS), which
is also called fools' gold, is common in all types of
rocks. It is the weathering of this mineral that leads to
acid-mine drainage which is nearly universal in coal
fields and metal sulfide mining regions.
9.7 Rocks
Three types of rock make up the crust of the earth.
Igneous rocks solidified from molten material either
within the earth (intrusive) or on or near the surface
(extrusive). Metamorphic rocks were originally
igneous or sedimentary rocks that were modified by
temperature, pressure, and chemically active fluids.
Sedimentary rocks are the result of the weathering of
preexisting rocks, erosion, and deposition. Geologists
have developed elaborate systems of nomenclature
and classification of rocks, but these are of little value
in hydrogeologic studies and therefore only the most
basic descriptions will be presented.
9.7.1 Igneous Rocks
Igneous rocks are classified on the basis of their
composition and grain size. Most consist of feldspar
and a variety of dark minerals; several others also
contain quartz. If the parent molten material cools
slowly deep below the surface, minerals will have an
opportunity to grow and the rock will be coarse
grained. Magma that cools rapidly, such as that
derived from volcanic activity, is so fine grained that
individual minerals generally cannot be seen even
with a hand lens. In some cases the molten material
began to cool slowly, allowing some minerals to grow,
and then the rate changed dramatically so that the
remainder formed a fine groundmass. This texture,
consisting of large crystals in a fine grained matrix, is
called porphyritic.
Intrusive igneous rocks can only be seen where they
have been exposed by erosion. They are concordant
if they more or less parallel the bedding of the
enclosing rocks and discordant if they cut across the
bedding. The largest discordant igneous masses are
called batholiths and they occur in the eroded centers
of many ancient mountains. Their dimensions are in
the range of tens of miles. Batholiths usually consist
largely of granite, which is surrounded by
metamorphic rocks.
Discordant igneous rocks include dikes ranging in
thickness from a few inches to thousands of feet.
Many are several miles long. Sills are concordant
bodies that have invaded sedimentary rocks along
bedding planes. They are relatively thin. Both sills and
dikes tend to cool quite rapidly and, resultingly, are
fine grained.
Extrusive rocks include lava flows or other types
associated with volcanic activity, such as the glassy
rock pumice, and the consolidated ash called tuff.
These are fine grained or even glassy.
With some exceptions, igneous rocks are dense and
have very little porosity or permeability. Most,
however, are fractured to some degree and can store
and transmit a modest amount ol water. Some lava
flows are notable exceptions because they contain
large diameter tubes or a permeable zone at the top
of the flow where gas bubbles migrated to the surface
before the rock solidified. These rocks are called
scoria.
9.7.2 Metamorphic Rocks
Metamorphism is a process that changes preexisting
rocks into new forms because of increases in
temperature, pressure, and chemically active fluids.
Metamorphism may affect igneous, sedimentary, or
other metamorphic rocks. The changes brought about
include the formation of new minerals, increase in
grain size, and modification of rock structure or
texture, all of which depend on the original rock's
composition and the intensity of the metamorphism.
Some of the most obvious changes are in texture,
which serves as a means of classifying metamorphic
rocks into two broad groups, the foliated and
nonfoliated rocks. Foliated metamorphic rocks typify
regions that have undergone severe deformation,
such as mountain ranges. Shale, which consists
mainly of silt and clay, is transformed into slate by the
change of clay to mica. Mica, being a platy mineral,
grows with its long axis perpendicular to the principle
direction of stress, forming a preferred orientation.
This orientation, such as the development of cleavage
in slate, may differ greatly from the original bedding.
With increasing degrees of metamorphism, the grains
of mica grow larger so that the rock has a distinct
foliation, which is characteristic of the metamorphic
rock schist. At even higher grades of metamorphism,
the mica may be transformed to a much coarser
grained feldspar, producing the strongly banded
texture of gneiss.
Nonfoliated rocks include the hornfels and another
group formed from rocks that consist mainly of a
single mineral. The hornfels occur around an intrusive
body and were changed by "baking" during intrusion.
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The second group includes marble and quartzite, as
well as several other forms. Marble is
metamorphosed limestone and quartzite is
metamorphosed quartz sandstone.
There are many different types of metamorphic rocks,
but from a hydrogeologic viewpoint they normally
neither store nor transmit much water and are of only
minor importance as aquifers. Their primary
permeability is notably small, if it exists at all, and
fluids are forced to migrate through secondary
openings, such as faults, joints, or other types of
fractures.
9.7.3 Sedimentary Rocks
Sedimentary rocks are deposited either in a body of
water or on the land by running water, by wind, and
by glaciers. Each depositional agent leaves a
characteristic stamp on the material it deposits. The
sediments carried by these agents were first derived
by the weathering and erosion of preexisting rocks.
The most common sedimentary rocks are shale,
siltstone, sandstone, limestone, and glacial till. The
change from a loose, unconsolidated sediment to a
rock is the process of lithification. Although
sedimentary rocks appear to be the dominant type, in
reality they make up but a small percentage of the
earth. They do, however, form a thin crust over much
of the earth's surface, are the type most readily
evident, and serve as the primary source of ground
water.
The major characteristics of sedimentary rocks are
sorting, rounding, and stratification. A sediment is well
sorted if the grains are nearly all the same size. Wind
is the most effective agent of sorting and this is
followed by water. Glacial till is unsorted and consists
of a wide mixture of material that ranges from large
boulders to clay.
While being transported, sedimentary material loses
its sharp, angular configuration as it develops some
degree of rounding. The amount of rounding depends
on the original shape, composition, transporting
medium, and the distance traveled.
Sorting and rounding are important features of both
consolidated and unconsolidated material because
they have a major control on permeability and
porosity. The greater the degree of sorting and
rounding, the higher will be the water-transmitting
and storage properties. This is why a deposit of sand,
in contrast to glacial till, can be such a productive
aquifer.
Most sedimentary rocks are deposited in a sequence
of layers or strata. Each layer or stratum is separated
by a bedding plane, which probably reflects variations
in sediment supply or some type of short term
erosion. Commonly bedding planes represent
changes in grain size. Stratification provides many
clues in our attempt to unravel geologic history. The
correlation of strata between wells or outcrops is
called stratigraphy.
Sedimentary rocks are classified on the basis of
texture (grain size and shape) and composition.
Clastic rocks consist of particles of broken or worn
material and include such shale, siltstone, sandstone,
and conglomerate. These rocks were lithified by
compaction, in the case of shale, and by
cementation. The most common cements are clay,
calcite, quartz, and limonite. The last three, carried by
ground water, precipitate in the unconsolidated
material under specific geochemical conditions.
The organic or chemical sedimentary rocks consist of
strata formed from or by organisms and by chemical
precipitates from sea water or other solutions. Most
have a crystalline texture. Some consist of well-
preserved organic remains, such as reef deposits and
coal seams. Chemical sediments include, in addition
to some limestones, the evaporites, such as halite
(sodium chloride), gypsum, and anhydrite. Anhydrite
is an anhydrous calcium sulfate.
Geologists also have developed an elaborate
classification of sedimentary rocks, which is of little
importance to the purpose of this introduction. In fact,
most sedimentary rocks are mixtures of clastic debris,
organic material, and chemical precipitates. One
should keep in mind not the various classifications,
but rather the texture, composition, and other features
that can be used to understand the origin and history
of the rock.
9.8 Weathering
Generally speaking, a rock is stable only in the
environment in which it was formed. Once removed
from that environment, it begins to change, rapidly in
some cases but more often slowly, by weathering.
The two major processes of weathering are
mechanical and chemical, but they usually proceed in
concert.
9.8.1 Mechanical Weathering
Mechanical weathering is the physical breakdown of
rocks and minerals. Some is the result of fracturing
due to the volumetric increase when water in a crack
turns to ice, some is the result of abrasion during
transport by water, ice, or wind, and a large part is
the result of gravity causing rocks to fall and shatter.
Mechanical weathering alone only reduces the size of
the rock; its chemical composition is not changed.
The weathered material formed ranges in size from
boulders to silt.
9.8.2 Chemical Weathering
Chemical weathering, on the other hand, is an actual
change in composition as minerals are modified from
one type to another. Many if not most of the changes
are accompanied by a volumetric increase or
decrease, which in itself further promotes additional
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chemical weathering. The rate depends on
temperature, surface area, and available water.
The major reactions involved in chemical weathering
are oxidation, hydrolysis, and carbonation. Oxidation
is a reaction with oxygen to form an oxide, hydrolysis
is reaction with water, and carbonation is a reaction
with C02 to form a carbonate. In these reactions the
total volume increases and, since chemical
weathering is most effective on grain surfaces,
disintegration occurs.
Quartz, whether vein deposits or individual grains,
undergoes practically no chemical weathering; the
end product is quartz sand. Some of the feldspars
weather to clay and release calcium, sodium, silica,
and many other elements that are transported in
water. The iron-bearing minerals provide, in addition
to iron and magnesium, weathering products that are
similar to the feldspars.
9.9 Erosion and Deposition
Once a rock begins to weather, the by-products
await erosion or transportation, which must be
followed by deposition. The major agents involved in
this part of the rock cycle are running water, wind,
and glacial ice.
9.9.7 Water borne Deposits
Mass wasting is the downslope movement of large
amounts of detrital material by gravity. Through this
process sediments are made available to streams that
carry them away to a temporary or permanent site of
deposition. During transportation some sorting occurs
and the finer silt and clay are carried farther
downstream. The streams, constantly filling, eroding,
and widening their channels, leave materials in their
valleys that indicate much of the history of the region.
Stream valley deposits, called alluvium, are shown on
geologic maps by the symbol Qal, meaning
Quaternary age alluvium. Alluvial deposits are distinct
but highly variable in grain size, composition, and
thickness. Where they consist of glacially derived
sand and gravel, called outwash, they form some of
the most productive water-bearing units in the world.
Sediments, either clastic or chemical/organic,
transported to past and present seas and ocean
basins spread out to form, after lithification, extensive
units of sandstone, siltstone, shale, and limestone. In
the geologic past, these marine deposits covered vast
areas and when uplifted they formed the land surface,
where they again began to weather in anticipation of
the next trip to the ocean.
The major features of marine sedimentary rocks are
their widespread occurrence and rather uniform
thickness and composition, although extreme
changes exist in many places. If not disturbed by
some type of earth movement, they are stratified and
horizontal. Furthermore, each lithologic type is unique
relative to adjacent units. The bedding planes or
contacts that divide them represent distinct
differences in texture or composition. From a
hydrologic perspective, differences in texture from
one rock type to another produce boundaries that
strongly influence ground-water flow. Consequently,
ground water tends to flow parallel to these
boundaries, that is, within a particular geologic
formation rather than across them.
9.9.2 Windborne Deposits
Wind-laid or eolian deposits are relatively rare in the
geologic record. The massively cross-bedded
sandstone of the Navajo Sandstone in Utah's Zion
National Park and surrounding areas is a classic
example in the United States. Other deposits are
more or less local and represent dunes formed along
beaches of large water bodies or streams. Their
major characteristic is the high degree of sorting.
Dunes, being relatively free of silt and clay, are very
permeable and porous, unless the openings have
been filled by cement. They allow rapid infiltration of
water and can form major water-bearing units, if the
topographic and geologic conditions are such that the
water does not rapidly drain.
Another wind-deposited sediment is loess, which
consists largely of silt. It lacks bedding but is typified
by vertical jointing. The silt is transported by wind
from deserts, flood plains, and glacial deposits. Loess
weathers to a fertile soil and is very porous. It is
common along the major rivers in the glaciated parts
of the United States and in China, parts of Europe,
and adjacent to deserts and deposits of glacial
outwash.
9.9.3 Glacial Deposits
Glaciers erode, transport, and deposit sediments that
range from clay to huge boulders. They subdue the
land surface over which they flow and bury former
river systems. The areas covered by glaciers during
the last Ice Age in the United States are shown in
Figure 9-12, but the deposits extend far beyond the
former margins of the ice. The two major types of
glaciers include valley or mountain glaciers and the
far more extensive continental glaciers. The deposits
they leave are similar, differing for the most part only
in scale.
As a glacier passes slowly over the land surface it
incorporates material from the underlying rocks into
the ice mass, only to deposit that material elsewhere
when the ice melts. During this process it modifies
the land surface, both through erosion and deposition.
The debris associated with glacial activity is
collectively termed glacial drift. Unstratified drift,
usually deposited directly by the ice, is glacial till, a
heterogeneous mixture of boulders, gravel, sand, silt,
and clay. Glacial debris reworked by streams and in
lakes is stratified drift. Although stratified drift may
range widely in grain size, the sorting far surpasses
198
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Figure 9-12 Area) extent of glacial deposits in the United States (from Heath, 1984).
Alaska
Areas occupied by lakes during the glacial period
••'•; •'•.,'! Areas underlain by glacial deposits
Figure 9-13 Dip and strike symbols commonly shown on geologic maps.
Map View
Anticline Syncline
Cross Section
The arrow indicates the direction of dip. In an anticline,
the rocks dip away from the crest and in a syncline they
dip toward the center.
199
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that of glacial till. Glacial lake clays are particularly
well sorted.
Glacial geologists usually map not on the basis of
texture but rather the type of landform that was
developed, such as moraines, outwash, drumlins, and
so on. The various kinds of moraines and associated
landforms are composed largely of unstratified drift
with incorporated layers of sand and gravel. Stratified
drift is found along existing or former stream valleys
or lakes that were either in the glacier or extended
downgradient from it. Meltwater stream deposits are
mixtures of sand and gravel. In places, some have
coalesced to develop extensive outwash plains.
Glaciers advanced and retreated many times,
reworking, overriding, and incorporating sediments
from previous advances into the ice, subsequently
redepositing them elsewhere. There was a constant
inversion of topography as buried ice melted causing
adjacent, waterlogged till to slump into the low areas.
During advances, the ice might have overridden older
outwash layers so that upon melting, these sand and
gravel deposits were covered by a younger layer of
till. Regardless of the cause, the final effect is one of
complexity of origin, history, and stratigraphy. When
working with glacial till deposits, it is nearly always
impossible to predict the lateral extent or thickness of
a particular lithology in the subsurface. Surficial
stratified drift is more uniform than till in the
thickness, extent, and texture.
9.10 Geologic Structure
A general law of geology is that in any sequence of
sedimentary rocks that has not been disturbed by
folding or faulting, the youngest unit is on the top. A
second general law is that sedimentary rocks are
deposited in a horizontal or nearly horizontal position.
The fact that rocks are found overturned, displaced
vertically or laterally, and squeezed into open or tight
folds clearly indicates that the crust of the earth is a
dynamic system. There is a constant battle between
the forces of destruction (erosion) and construction
(earth movements).
An unconformity is a break in the geologic record. It
is caused by a cessation in deposition that is followed
by erosion and subsequent deposition. The geologic
record is lost by the period of erosion because the
rocks that contained the record were removed.
If a sequence of strata is horizontal but the contact
between two rock groups in the sequence represents
an erosional surface, that surface is said to be a
disconformity. Where a sequence of strata has been
tilted and eroded and then younger, horizontal rocks
are deposited over them, the contact is an angular
unconformity. A nonconformity occurs where eroded
igneous or metamorphic rocks are overlain by
sedimentary rocks.
9.10.1 Folding
Rocks folded by compressional forces are common in
and adjacent to former or existing mountain ranges
(Figure 9-13). The folds range from a few inches to
50 miles or so across. Anticlines are rocks folded
upward into an arch. Their counterpart, synclines, are
folded downward like a valley. A monocline is a
flecture in which the rocks are horizontal, or nearly
so, on either side of the flecture.
Although many rocks have been folded into various
structures, this does not mean that these same
structures form similar topographic features. As the
folding takes place over eons, the forces of erosion
attempt to maintain a low profile. As uplift continues,
erosion removes weathering products from the rising
mass, carrying them to other places of deposition.
The final topography is related to the erodibility of the
rocks, with resistant strata such as sandstone forming
ridges, and the less resistant material such as shale
forming valleys. Consequently, the geologic structure
of an area may bear little resemblance to its
topography.
The structure of an area can be determined from field
studies or a geologic map, if one exists. Various types
of folds and their dimensions appear as unusual
patterns on geologic maps. An anticline, for example,
will be depicted as a series of rock units in which the
oldest is in the middle, while a syncline is represented
by the youngest rock in the center (Figure 9-13).
More or less equidimensional anticlines and synclines
are termed domes and basins, respectively.
The inclination of the top of a fold is the plunge. Folds
may be symmetrical, asymmetrical, overturned, or
recumbent. The inclination of the rocks is indicated
by dip and strike symbols. The strike is perpendicular
to the dip and the degree of dip is commonly shown
by a number (Figure 9-13). The dip may range from
less than a degree to vertical.
9.70.2 Fractures
Fractures in rocks are either joints or faults. A joint is
a fracture along which no movement has taken place;
a fault implies movement. Movement along faults is
as little as a few inches to tens of miles. Probably all
consolidated rocks and a good share of the
unconsolidated deposits contain joints. Although not
well recognized by most individuals involved in
ground-water problems, joints exert a major control
on water movement and chemical quantity.
Characteristically joints are open and serve as major
conduits or pipes. Water can move through them
quickly, perhaps carrying contaminants, and, being
open, the filtration effect is lost. It is a good possibility
that the outbreak of many waterborne diseases that
can be traced to ground-water supplies are the
result of the transmission of infectious agents through
fractures to wells and springs.
200
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Faults are most common in the deformed rocks of
mountain ranges, suggesting either lengthening or
shortening of the crust. Movement along a fault may
be horizontal, vertical, or a combination. The most
common types of faults are called normal, reverse,
and lateral (Figure 9-14). A normal fault, which
indicates stretching of the crust, is one in which the
upper or hanging wall has moved down relative to the
lower or foot wall. The Red Sea, Dead Sea, and the
large lake basins in the east African highlands, among
many others, lie in a graben, which is a block
bounded by normal faults. A reverse or thrust fault
implies compression and shortening of the crust. It is
distinguished by the fact that the hanging wall has
moved up relative to the foot wall. A lateral fault is
one in which the movement has been largely
horizontal. The San Andreas Fault, extending some
600 miles from San Francisco Bay to the Gulf of
California, is the most notable lateral fault in the
United States. It was movement along this fault that
produced the 1906 San Francisco earthquake.
9.11 Geologic Time
Geologic time deals with the relation between the
emplacement or disturbance of rocks and time. The
geologic time scale was developed in order to provide
some standard classification (Table 9-3). It is based
on a sequence of rocks that were deposited during a
particular time interval. The divisions are commonly
based on some type of unconformity. In considering
geologic time, three types of units are defined. They
are rock units, time and rock units, and time units.
9.11.1 Rock Units
A rock unit refers to some particular lithology. These
may be further divided into geologic formations which
are of sufficient size and uniformity to be mapped in
the field. The Pierre Shale, for example, is a
widespread and, in places, thick geologic formation
that extends over much of the Northern Great Plains.
Formations can also be divided into smaller units
called members. Formations have a geographic name
that may be coupled with a term that describes the
major rock type. Two or more formations comprise a
group.
Table 9-3 Geologic Time Scale
Era
Cenozoic
Mesozoic
Paleozoic
Precambrian
Period
Quaternary
Tertiary
Cretaceous
Jurassic
Triassic
Permian
Pennsylvanian
Mississippian
Devonian
Silurian
Ordovician
Cambrian
Lasted at least 2.
Epoch
Recent
Pleistocene
Pliocene
Miocene
Oligocene
Eocene
Paleocene
5 billion years
Millions of
Years Ago
0-1
1-13
13-25
25-36
36-58
58-63
63-135
135-181
181-230
230-280
280-310
310-345
345-405
405-425
425-500
500-600
9.11,2 Time and Time-Rock Units
Time-rock units refer to the rock that was deposited
during a certain period of time. These units are
divided into system, series, and stage. Time units
refer to the time during which a sequence of rocks
was deposited. The time-rock term "system" has
the equivalent time term, "period." That is, during the
Cretaceous Period, for example, rocks of the
Cretaceous System were deposited, consisting of
many groups and formations. Time units are named
in such a way that the eras reflect the complexity of
life forms that existed, such as the Mesozoic or
"middle life." System or period nomenclature is
largely based on the geographic location in which the
rocks were first described, such as Jurassic, which
relates to the Jura Mountains of Europe.
The terms used by geologists to describe rocks
relative to geologic time are useful to the ground-
water investigator in that they allow one to better
perceive a regional geologic situation. The terms
alone have no significance as far as water-bearing
properties are concerned.
Figure 9-14 Cross sections of normal, reverse, and lateral faults.
-\ /-Foot Wall
Fault
Cross-Section
of
Normal Fault
Hanging Wall
Hanging Wall
Foot Wall -
Cross-Section
of
Reverse Fault
Normal Fault -
V
Graben
Plan View
of
Lateral Fault
Cross-Section
of
Graben
Normal Fault
201
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9.12 References
Blatt, H., G. Middleton, and R. Murray. 1980. Origin
of Sedimentary Rocks. 2nd ed. Prentice-Hall
Publishing Co., Inc., Englewood Cliffs, NJ.
Ernst, W.G. 1969. Earth Materials. Prentice-Hall
Publishing Co., Inc., Englewood Cliffs, NJ.
Flint, R.F. 1971. Glacial and Quaternary Geology.
John Wiley & Sons, New York, NY.
Foster, R.J. 1971. Geology. Charles E. Merrill
Publishing Co., Columbus, OH.
Heath, R.C. 1984. Ground-Water Regions of the
United States. U.S. Geological Survey Water-Supply
Paper 2242, U.S. Government Printing Office,
Washington, DC.
Pettyjohn, W.A., J.R.J. Studlick, and R.C. Bain. 1979.
Quality of Drinking Water in Rural America. Water
Technology 7/8.
Sawkins, F.J., C.G. Chase, D.G. Darby, and G. Rapp,
Jr. 1978. The Evolving Earth, A Text in Physical
Geology. Macmillan Publishing Co., Inc., New York,
NY.
Spencer, E.W. 1977. Introduction to the Structure of
the Earth. 2nd ed. McGraw-Hill Book Co., Inc., New
York, NY.
Tarbuck, E.J., and F.K. Lutgens. 1984. The Earth, An
Introduction to Physical Geology. Charles E. Merrill
Publishing Co., Inc., Columbus, OH.
Tolman, C.F. 1937. Ground Water. McGraw-Hill
Book Co., Inc., New York, NY.
202
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APPENDIX
SOURCES OF INFORMATION ABOUT GROUND-WATER CONTAMINATION
INVESTIGATIONS
SOLID AND HAZARDOUS WASTE
AGENCIES
ALABAMA
Daniel E. Cooper, Director
Land Division
Alabama Dept. of Environmental Management
1751 Federal Drive
Montgomery, AL 36130
Phone: (205) 271-7730
ALASKA
Stan Hungerford
Air and Solid Waste Management
Dept. of Environmental Conservation
Pouch O
Juneau, AK 99811
Phone: (907) 465-2635
AMERICAN SAMOA
Pati Faiai, Executive Secretary
Environmental Quality Commission
American Samoa Government
Pago Pago, American Samoa 96799
Phone: Overseas Operator 633-4116
Randy Morris, Deputy Director
Department of Public Works
Pago Pago, American Samoa 96799
ARIZONA
Ron Miller, Manager
Office of Waste and Water Quality Management
Arizona Dept. of Health Services
2005 North Central Avenue
Phoenix, AZ 85004
Phone: (602) 257-2305
ARKANSAS
Vincent Blubaugh, Chief
Solid & Hazardous Waste Division
Dept. of Pollution Control and Ecology
P.O. Box 9583
8001 National Drive
Little Rock, AR 72219
Phone: (501) 562-7444
CALIFORNIA
Vacant, Deputy Director
Toxic Substances Control Programs
Dept. of Health Services
714 P Street, Room 1253
Sacramento, CA 95814
Phone: (916) 322-7202
Michael Campos, Executive Director
State Water Resources Control Board
P.O. Box 100
Sacramento, CA 95801
Phone: (916) 445-1553
Sherman E. Roodzant, Chairman
California Waste Management Board
1020 Ninth Street, Suite 300
Sacramento, CA 95814
Phone: (916) 322-3330
COLORADO
Kenneth Waesche, Director
Waste Management Division
Colorado Dept. of Health
4210 E. 11th Avenue
Denver, CO 80220
Phone: (303) 320-8333
COMMONWEALTH OF NORTHERN MARIANA
ISLANDS
George Chan, Administrator
Division of Environmental Quality
Dept. of Public Health and Environmental Services
Commonwealth of the Northern Mariana Islands
Saipan, CM 96950
Phone: Overseas Operator-6984
CONNECTICUT
Stephen Hitchock, Director
Hazardous Material Management Unit
Dept. of Environmental Protection
State Office Building
165 Capitol Avenue
Hartford, CT06106
Phone: (203) 566-4924
203
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Michael Cawley
Connecticut Resource Recovery Authority
179 Allyn Street, Suite 603
Professional Building
Hartford, CT06103
Phone: (203) 549-6390
DELAWARE
William Razor, Supervisor
Solid Waste Management Branch
Dept. of Natural Resources and Environmental
Control
89 Kings Highway
P.O. Box 1401
Dover, DE 19901
Phone: (302) 736-4781
DISTRICT OF COLUMBIA
Angelo Tompros, Chief
Dept. of Consumer & Regulatory Affairs
Pesticides & Hazardous Waste Management
Room 112
5010 Overlook Avenue, S.W.
Washington, DC 20032
Phone: (202) 767-8422
FLORIDA
Robert W. McVety, Administrator
Solid & Hazardous Waste Section
Dept. of Environmental Regulation
Twin Towers Office Building
2600 Blair Stone Road
Tallahassee, FL 32301
Phone: (904) 488-0300
GEORGIA
John Taylor, Chief
Land Protection Branch
Environmental Protection Division
Dept. of Natural Resources
270 Washington Street, S.W., Room 723
Atlanta, GA 30334
Phone: (404) 656-2833
GUAM
James Branch, Administrator
Guam Environmental Protection Agency
P.O. Box 2999
Agana, GU 96910
Phone: Overseas Operator 646-8863
HAWAII
Melvin Koziumi, Deputy Director
Environmental Health Division
Dept. of Health
P.O. Box 3378
Honolulu, HI 96801
Phone: (808) 548-4139
IDAHO
Steve Provant, Supervisor
Hazardous Materials Bureau
Dept. of Health & Welfare
State House
Boise, ID 83720
Phone: (208) 334-2293
ILLINOIS
Robert Kuykendall, Manager
Division of Land Pollution Control
Environmental Protection Agency
2200 Churchill Road, Room A-104
Springfield, IL 62706
Phone: (217) 782-6760
William Child, Deputy Manager
Division of Land Pollution Control
Environmental Protection Agency
2200 Churchill Road, Room A-104
Springfield, IL 62706
Phone: (217) 782-6760
INDIANA
David Lamm, Director
Land Pollution Control Division
State Board of Health
1330 West Michigan Street
Indianapolis, IN 46206
Phone: (317) 633-0619
IOWA
Ronald Kolpa
Hazardous Waste Program Coordinator
Dept. of Water, Air & Waste Management
Henry A. Wallace Building
900 East Grand
Des Moines, IA50319
Phone: (515) 281-8925
KANSAS
Dennis Murphey, Manager
Bureau of Waste Management
Dept. of Health & Environment
Forbes Field, Building 321
Topeka, KS 66620
Phone: (913) 862-9360
KENTUCKY
J. Alex Barber, Director
Division of Waste Management
Dept. of Environmental Protection
Cabinet for Natural Resources and Environmental
Protection
18 Reilly Road
Frankfort, KY 40601
Phone: (502) 564-6716
204
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LOUISIANA
Gerald J. Healy, Administrator
Solid Waste Management Division
Dept. of Environmental Quality
P.O. Box 44307
Baton Rouge, LA 70804
Phone: (504) 342-1216
Glenn Miller, Administrator
Hazardous Waste Management Division
Dept. of Environmental Quality
P.O. Box 44307
Baton Rouge, LA 70804
Phone: (504) 342-1227
MAINE
David Boulter, Director
Licensing and Enforcement Division
Bureau of Oil & Hazardous Materials
Dept. of Environmental Protection
State House ~ Station 17
August, ME 04333
Phone: (207) 289-2651
MARYLAND
Bernard Bigham
Waste Management Administration
Dept. of Health & Mental Hygiene
201 W. Preston Street, Room 212
Baltimore, MD 21201
Phone: (301) 225-5649
Alvin Bowles, Chief
Hazardous Waste Division
Waste Management Administration
Dept. of Health & Mental Hygiene
201 W. Preston Street, Room 212
Baltimore, MD 21201
Ronald Nelson, Director
Waste Management Administration
Office of Environmental Programs
Dept. of Health & Mental Hygiene
201 W. Preston Street, Room 212
Baltimore, MD 21201
Phone: (301) 225-5647
MASSACHUSETTS
William Cass, Director
Division of Solid & Hazardous Waste
Dept. of Environmental Quality
Engineering
One Winter Street
Boston, MA 02108
Phone: (617) 292-5589
MICHIGAN
Delbert Rector, Chief
Hazardous Waste Division
Environmental Protection Bureau
Dept. of Natural Resources
Box 30028
Lansing, Ml 48909
Phone: (517) 373-2730
Allan Howard, Chief
Technical Services Section
Hazardous Waste Division
Dept. of Natural Resources
Box 30028
Lansing, Ml 48909
Phone: (517) 373-8448
MINNESOTA
Dale L. Wikre, Director
Solid and Hazardous Waste Division
Pollution Control Agency
1935 West County Road, B-2
Roseville, MN55113
Phone: (612) 296-7282
MISSISSIPPI
Jack M. McMillan, Director
Division of Solid & Hazardous Waste Management
Bureau of Pollution Control
Dept. of Natural Resources
P.O. Box 10385
Jackson, MS 39209
Phone: (601) 961-5062
MISSOURI
Dr. David Bedan, Director
Waste Management Program
Dept. of Natural Resources
117 East Dunklin Street
P.O. Box 176
Jefferson City, MO 65102
Phone: (314) 751-3241
MONTANA
Duane L. Robertson, Chief
Solid Waste Management Bureau
Dept. of Health and Environmental Sciences
Cogswell Building
Helena, MT 59602
Phone: (406) 444-2821
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NEBRASKA
Mike Steffensmeier
Section Supervisor
Hazardous Waste Management Section
Dept. of Environmental Control
State House Station
P.O. Box 94877
Lincoln, NE 68509
Phone: (402) 471-2186
NEVADA
Verne Rosse
Waste Management Program Director
Division of Environmental Protection
Dept. of Conservation and Natural Resources
Capitol Complex
201 South Fall Street
Carson City, NV 89710
NEW HAMPSHIRE
Dr. Brian Strohm, Assistant Director
Division of Public Health Services
Office of Waste Management
Dept. of Health and Welfare
Health and Welfare Building
Hazen Drive
Concord, NH 03301
Phone: (603) 271-4608
NEW JERSEY
Dr. Man/van Sadat, Director
Division of Waste Management
Dept. of Environmental Protection
32 E. Hanover Street, CN-027
Trenton, NJ 08625
Phone: (609) 292-1250
NEW MEXICO
Richard Perkins, Acting Chief
Groundwater & Hazardous Waste Bureau
Environmental Improvement Division
New Mexico Health & Environment Dept.
P.O. Box 968
Santa Fe, NM 87504-0968
Phone: (505) 984-0020
Peter Pache, Program Manager
Hazardous Waste Section
Groundwater & Hazardous Waste Bureau
Environmental Improvement Division
New Mexico Health & Environment Dept.
P.O. Box 968
Santa Fe, NM 87504-0968
Phone: (505) 984-0020
NEW YORK
Norman H. Nosenchuck, Director
Division of Solid & Hazardous Waste
Dept. of Environmental Conservation
50 Wolf Road, Room 209
Albany, NY 12233
Phone: (518) 457-6603
NORTH CAROLINA
William L. Meyer, Head
Solid & Hazardous Waste Management Branch
Division of Health Services
Dept. of Human Resources
P.O. Box 2091
Raleigh, NC 27602
Phone: (919) 733-2178
NORTH DAKOTA
Martin Schock, Director
Division of Hazardous Waste
Management and Special Studies
Dept. of Health
1200 Missouri Avenue, 3rd Floor
Bismarck, ND 58501
Phone: (701) 224-2366
OHIO
Steven White, Chief
Division of Solid & Hazardous Waste Management
Ohio Environmental Protection Agency
361 East Broad Street
Columbus, OH 43215
Phone: (614) 466-7220
OKLAHOMA
Dwain Farley, Chief
Waste Management Service
Oklahoma State Dept. of Health
P.O. Box 53551
Oklahoma City, OK 73152
Phone: (405) 271-5338
OREGON
Mike Downs, Administrator
Hazardous & Solid Waste Division
Dept. of Environmental Quality
P.O. Box 1760
Portland, OR 97207
Phone: (503) 229-5356
PENNSYLVANIA
Donald A. Lazarchik, Director
Bureau of Solid Waste Management
Dept. of Environmental Resources
Bulton Building, 8th Floor
P.O. Box 2063
Harrisburg, PA 17120
Phone: (717) 787-9870
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PUERTO RICO
Santos Rohena, Director
Solid, Toxics, & Hazardous Waste Program
Environmental Quality Board
P.O. Box 11488
Santurce, PR 00910-1488
Phone: (809) 725-0439
RHODE ISLAND
John S. Quinn, Jr., Chief
Solid Waste Management Program
Dept. of Environmental Management
204 Cannon Building
75 Davis Street
Providence, Rl 02908
Phone: (401) 277-2797
SOUTH CAROLINA
Robert E. Malpass, Chief
Bureau of Solid and Hazardous Waste Management
South Carolina Dept. of Health & Environmental
Control
2600 Bull Street
Columbia, SC 29201
Phone: (803) 758-5681
SOUTH DAKOTA
Joel C. Smith, Administrator
Office of Air Quality & Solid Waste
Dept. of Water & Natural Resources
Joe Foss Building
Pierre, SD 57501
Phone: (605) 773-3329
TENNESSEE
Tom Tiesler, Director
Division of Solid Waste Management
Bureau of Environmental Services
Tennessee Dept. of Public Health
150 9th Avenue, North
Nashville, TN 37203
Phone: (615) 741-3424
TEXAS
Jack Carmichael, Chief
Bureau of Solid Waste Management
Texas Dept. of Health
1100 West 49th Street, T-602
Austin, TX 78756-3199
Phone: (512) 458-7271
Jay Snow, Chief
Solid Waste Section
Texas Dept. of Water Resources
1700 North Congress
P.O. Box 13087, Capitol Station
Austin, TX 78711
Phone: (512) 463-8177
UTAH
Dale Parker, Director
Bureau of Solid & Hazardous Waste Management
Dept. of Health
P.O. Box 2500
150 West North Temple
Salt Lake City, UT84110
Phone: (801) 533-4145
VERMONT
Richard A. Valentinetti, Director
Air and Solid Waste Programs
Agency of Environmental Conservation
State Office Building
P.O. Box 489
Montpelier, VT 05602
Phone: (802) 828-3395
VIRGIN ISLANDS
Robert V. Eepoel, Director
Hazardous Waste Program
Division of Natural Resources
Dept. of Conservation and Cultural Affairs
P.O. Box 4340, Charlotte Amalie
St. Thomas, VI 00801
Phone: (809) 774-6420
VIRGINIA
William F. Gilley, Director
Division of Solid and Hazardous Waste Management
Virginia Dept. of Health
Monroe Building, 11th Floor
101 North 14th Street
Richmond, VA23219
Phone: (804) 225-2667
WASHINGTON
Warl Tower, Supervisor
Solid & Hazardous Waste Management Division
Dept. of Ecology
Olympia, WA 98504
Phone: (296) 459-6316
Linda L. Brothers, Assistant Director
Office of Hazardous Substance & Air Quality Program
Dept. of Ecology
Olympia, WA 98504
Phone: (296) 459-6253
WEST VIRGINIA
Timothy Larway, Chief
Division of Water Resources
Dept. of Natural Resources
1201 Greenbrier Street
Charleston, WV 25311
Phone: (304) 348-5935
207
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WISCONSIN
Paul Didier, Director
Bureau of Solid Waste Management
Dept. of Natural Resources
P.O. Box 7921
Madison, Wl 53707
Phone: (608) 266-1327
WYOMING
Charles Porter, Supervisor
Solid Waste Management Program
State of Wyoming
Dept. of Environmental Quality
Equality State Bank Building
401 West 19th Street
Cheyenne, WY 82002
Phone: (307) 777-7752
208
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U.S. EPA OFFICE
PROTECTION
Ms. Marian Mlay, Diretor
Office of Ground-Water
U.S. EPA
401 M Street, SW
Washington, DC 20460
Phone: (202) 382-7077
Ms. Carol Wood, Director
Office of Ground-Water
Water Management Division
U.S. EPA, Region I
JFK Federal Building (Room 2113)
Boston, MA 02203
Phone: (617) 223-6486
Mr. John Malleck, Director
Office of Ground-Water
Water Management Division
U.S. EPA, Region II
26 Federal Plaza (Room 805)
New York, NY 10278
Phone: (212) 264-5635
Mr. Tom Merski, Director
Office of Ground-Water
Water Management Division
U.S. EPA, Region III
841 Chestnut Street
Philadelphia, PA 19107
Phone: (215) 597-2786
Mr. Stallings Howell, Director
Office of Ground-Water
Water Management Division
U.S. EPA, Region IV
345 Courtland Street, N.E.
Atlanta, GA 30365
Phone: (404) 881-7731
Mr. Charles Job, Acting Director
Office of Ground-Water
Water Management Division
U.S. EPA, Region V
230 S. Dearborn Street
Chicago, IL 61604
Phone: (312) 353-2406
Mr. Ken Kirkpatrick, Acting Director
Office of Ground-Water
Water Management Division
U.S. EPA, Region VI
1201 Elm Street
Dallas, TX 75270
Phone: (214) 767-2656
OF GROUND-WATER
Protection (WH-550G)
Mr. Timothy Amsden, Director
Office of Ground-Water
Water Management Division
U.S. EPA, Region VII
726 Minnesota Avenue
Kansas City, KS 66101
Phone: (913) 236-2815
Mr. Richard Long, Director
Water Management Division
U.S. EPA, Region VIII
999 18th Street
Denver, CO 80295
Phone: (303) 293-1543
Mr. James Thompson, Director
Office of Ground-Water
Water Management Division
U.S. EPA, Region IX
215 Fremont Street
San Francisco, CA 94105
Phone: (415) 974-8267
Mr. William A. Mullen, Director
Office of Ground-Water
Water Management Division
U.S. EPA, Region X (M/S 437)
1200 6th Avenue
Seattle, WA98101
Phone: (206) 442-1216
209
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FEDERAL INTERAGENCY GROUND-
WATER PROTECTION COMMITTEE
DEPARTMENT OF THE INTERIOR
Mr. Wayne N. Merchant
(Principal Agency contact)
Acting Assistant Secretary for Water and Science
U.S. Dept. of the Interior
18th &C Street, NW
Washington, DC 20240
Phone: (202) 343-2186
Mr. Harold W. Furman III
(EPA/Off. Gr. Wtr. Prot. Contact)
Deputy Assistant Secretary for Water and Science
U.S. Dept. of the Interior (Room 6652)
18th &C Street, NW
Washington, DC 20240
Phone: (202) 343-4811
Mr. Roland Dolly
(Bureau of Reclamation representative)
Special Assistant to the Commissioner
Bureau of Reclamation
U.S. Dept. of the Interior (Room 7641)
18th &C Street, NW
Washington, DC 20240
Phone: (202) 343-4115
Mr. Robert Kleinmann
(Bureau of Mines representative)
Pittsburg Research Center
Bureau of Mines
U.S. Dept. of the Interior
Cochrans Mill Road
P.O. Box 18070
Pittsburgh, PA 15236
Phone: (412) 675-6555
Mr. Phillip Cohen
(U.S. Geological Survey representative)
Chief Hydrologist
U.S. Geological Survey
U.S. Dept. of the Interior
409 National Center
Reston, VA 22092
Mr. William Horn
(Principal Agency contact)
Assistant Secretary for Fish, Wildlife and Parks
U.S. Dept. of the Interior
18th &C Street, NW
Washington, DC 20240
Phone: (202) 343-4416
Mr. Donald S. Herring
(EPA/Off. Gr. Wtr. Prot. contact)
Engineering & Safety Service Division
National Park Service (610)
U.S. Dept. of the Interior
P.O. Box 37127
Washington, DC 20013
Phone: (202) 343-7040
Mr. Dan Kimball
(National Park Service representative)
Water Resources Division
National Park Service - Air
U.S. Dept. of the Interior (Room 7641)
P.O. Box 25287
Denver, CO 80225
Phone: (202) 776-8765
Mr. Hal O'Conner
(Fish & Wildlife Service representative)
Associate Director
Habitat Resources
Fish & Wildlife Service
U.S. Dept. of the Interior
18th &C Street, NW
Washington, DC 20240
Phone: (202) 343-4767
Mr. Steve Griles
(Principal Agency contact)
Acting Assistant Secretary for Lands & Mineral
Management
U.S. Dept. of the Interior
18th & C Street, NW
Washington, DC 20240
Phone: (202) 343-2186
Mr. Dan Muller
(Bureau of Land Management representative)
Bureau of Land Management
U.S. Dept. of the Interior
Premier Building (W0222)
18th & C Street, NW
Washington, DC 20240
Phone: (202) 653-9210
Mr. Al Perry
(Office of Surface Mining representative)
Office of Surface Mining
U.S. Dept. of the Interior
18th &C Street, NW
Washington, DC 20240
Phone: (202) 343-5854
210
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DEPARTMENT OF AGRICULTURE
Mr. George Dunlap
(Principal Agency contact)
Assistant Secretary for Natural Resources and
Environment
U.S. Dept. of Agriculture
Administration Building (217E)
Washington, DC 20250
Phone: (202) 447-7173
Mr. John Vance
(EPA/Off. Gr. Wtr. Prot. contact)
Forest Service
U.S. Dept. of Agriculture
Room 4207
Box 2417
Washington, DC 20013
Phone: (202) 447-7947
Mr. Louis Kirkaldie
(Soil Conservation Service representative)
Soil Conservation Service
U.S. Dept. of Agriculture
Room 6132
P.O. Box 2890
Washington, DC 20013
Phone: (202) 447-5858
Mr. Fred Swader
(Extension Service representative)
Extension Service
U.S. Dept. of Agriculture
Room 3340
South Agriculture Building
14th & Independence, SW
Washington, DC 20250
Phone: (202) 447-5369
DEPARTMENT OF JUSTICE
Mr. F. Henry Habicht II
(Principal Agency contact)
Assistant Attorney General
Land and Natural Resources
U.S. Dept. of Justice
10th Street & Constitution Avenue, NW
Washington, DC 20530
Phone: (202) 633-2701
Mr. Myles E. Flint
(EPA/Off. Gr. Wtr. Prot. contact)
Deputy Assistant Attorney General
Land and Natural Resources Division
U.S. Dept. of Justice
10th Street & Constitution Avenue, NW
Washington, DC 20530
Phone: (202) 633-2718
Mr. David Buenta
(representative)
Section Chief
Environmental Enforcement Section
Land and Natural Resources Division
U.S. Dept. of Justice
10th Street & Constitution Avenue, NW
Washington, DC 20530
Phone: (202) 633-5271
Ms. Marcy Toney
(representative)
Attorney
Policy, Legislation & Special Litigation Section
U.S. Dept. of Justice
Room 2613, Main Justice
10th Street & Constitution Avenue, NW
Washington, DC 20530
Phone: (202) 633-1442
DEPARTMENT OF THE ARMY
LTG E.R. Heiberg, III
(Principal Agency contact)
Commander U.S. Army
Corps of Engineers
U.S. Dept. of the Army
Pulaski Building
20 Massachusetts Avenue, NW
Washington, DC 20314
Phone: (202) 272-0000
Mr. Ming T. Tseng
(EPA/Off. Gr. Wtr. Prot. contact)
Office of Chief Engineer
U.S. Army Corps of Engineers
U.S. Dept. of the Army (DAEN-CWH-W)
Pulaski Building
20 Massachusetts Avenue, NW
Washington, DC 20314
Phone: (202) 272-8511
DEPARTMENT OF ENERGY
Mr. William A. Vaughan
(Principal Agency contact)
Assistant Secretary for Environment, Safety & Health
U.S. Dept. of Energy
Forrestal Building
1000 Independence Avenue, SW
Washington, DC 20585
Phone: (202) 252-4700
Mr. Ted Williams
(EPA/Off. Gr. Wtr. Prot. contact)
Director
Office of Policy Planning and Analysis
U.S. Dept. of Energy
Forrestal Building
1000 Independence Avenue, SW
Washington, DC 20585
Phone: (202) 252-2061
211
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Dr. Robert J. Stern
(Off. Env. Compliance representative)
Director
Office of Environmental Guidance
U.S. Dept. of Energy
Forrestal Building
1000 Independence Avenue, SW
Washington, DC 20585
Phone: (202) 252-4600
DEPARTMENT OF TRANSPORTATION
Mr. Ray A. Barnhart
(Principal Agency contact)
Administrator, Federal Highway Administration
U.S. Dept. of Transportation
Nassif Building
400 7th Street, SW
Washington, DC 20590
Phone: (202) 426-0650
Mr. Charles R. DesJardins
(Off. Env. Compliance representative)
Ecologist
Office of Environmental Policy (HEV-20)
Federal Highway Administration
U.S. Dept. of Transportation
Nassif Building
400 7th Street, SW
Washington, DC 20590
Phone: (202) 426-9173
DEPARTMENT OF DEFENSE
Mr. Carl J. Schafer, Jr.
(Principal Agency contact)
Director, Environmental Policy
U.S. Dept. of Defense
Room 3D 833
Pentagon
Washington, DC 20301
Phone: (202) 685-7820
Mr. Andres Talts, P.E.
(Def. Environ. Leader. Proj. representative)
Acting Director
Defense Environmental Leadership Project
U.S. Dept. of Defense
Room 202
1717 H Street, NW
Washington, DC 20006
Phone: (202) 653-1273
TENNESSEE VALLEY AUTHORITY
Honorable Charles H. Dean, Jr.
(Principal Agency contact)
Chairman
Tennessee Valley Authority
TVA Building
400 West Summit Hill Drive
Knoxville, TN 37902
Phone: (615) 632-2101
Mr. Robert Johnson
(EPA/Off. Gr. Wtr. Prot. contact)
Tennessee Valley Authority
215 Summer Place
Building 309
Walnut Street
Knoxville, TN 37902
Phone: (615) 632-6599
DEPARTMENT OF HEALTH & HUMAN SERVICES
Dr. James Mason
(Dept. Health & Human Serv. representative)
Acting Assistant Secretary, Public Health Service)
U.S. Dept. of Health & Human Services
Hubert H. Humphrey Building
200 Independence Avenue, SW
Washington, DC 20201
Phone: (202) 245-7694
Dr. Henry Falk
(Centers for Disease Control representative)
Centers for Disease Control
Center for Environmental Health
1600 Clifton Road
Atlanta, GA 30333
Phone: (404) 236-4095
NUCLEAR REGULATORY COMMISSION
Dr. Malcolm R. Knapp
(Principal Agency contact)
Chief, Geotechnical Branch
Division of Waste Management
Office of Nuclear Material Safety & Safeguards
U.S. Nuclear Regulatory Commission (M/S 623-SS)
Washington, DC 20555
Phone: (301) 427-4411
Mr. Michael Weber
(Nuclear Regulatory Commission representative)
U.S. Nuclear Regulatory Commission (M/S 623-SS)
Washington, DC 20555
Phone: (301) 427-4746
NATIONAL SCIENCE FOUNDATION
Mr. Nam P. Suh
(National Science Foundation representative)
Assistant Director for Engineering
National Science Foundation
1800 G Street, NW
Washington, DC 20550
Phone: (202) 357-7737
Mr. Edward H. Bryan
(National Science Foundation representative)
Program Director, Environmental Engineering
National Science Foundation
1800 G Street, NW
Washington, DC 20550
Phone: (202) 357-7737
212
•fl U . S . GOVERNMENT PRINTING OFFICE! 1987-748-121/40703
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